Hayes' Handbook of Pesticide Toxicology, Two-Volume Set, Third Edition volume 1 & 2

Hayes’ Handbook of Pesticide Toxicology

(c) 2011 Elsevier Inc. All Rights Reserved.

Hayes’ Handbook of Pesticide Toxicology Third Edition VOLUME 1 & 2

EDITED BY Robert Krieger University of California, Riverside

ASSOCIATE EDITORS: John Doull Joop van Hemmen† Ernest Hodgson Howard Maibach Lawrence Reiter Leonard Ritter John Ross William Slikker EDITORIAL ASSISTANT: Helen Vega †

Deceased

AMSTERDAM • BOSTON • HEIDELBERG • LONDON • NEW YORK • OXFORD • PARIS SAN DIEGO • SAN FRANCISCO • SINGAPORE • SYDNEY • TOKYO Academic Press is an imprint of Elsevier

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Academic Press is an imprint of Elsevier 32 Jamestown Road, London NW1 7BY, UK 30 Corporate Drive, Suite 400, Burlington, MA 01803, USA 525 B Street, Suite 1800, San Diego, CA 92101-4495, USA First edition 1991 Second edition 2001 Third edition 2010 Copyright © 1991, 2001, 2010 Elsevier Inc. All rights reserved with the exception of Chapter 2 © 2009 American Chemical Society. Chapters 18, 34 and 61 in the Public Domain. Chapters 31, 49, 58, 63, 72, 78, 82, 86, 91, 92 and 107 © 2001 Elsevier Inc. All rights reserved.

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10 9 8 7 6 5 4 3 2 1

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Dedication

Wayland Jackson “Jack” Hayes, Jr. made enduring contributions to pesticide science. Hayes’ Handbook of Pesticide Toxicology, third edition, carries his name to recognize his profound commitment to “improve the knowledge of toxicology, in general, the epidemiology of pesticide poisoning, and the medical management of cases.” He wrote and spoke often of the importance of the first principles of toxicology as Chief Toxicologist at Centers for Disease Control, Atlanta, Georgia, and later as Professor of Toxicology, School of Medicine, Vanderbilt University, Nashville, Tennessee. Hayes contributed his first volume to the toxicological literature as the Clinical Handbook on Economic Poisons (1963), replacing “Clinical Memoranda on Economic Poisons” first issued in March 1950 as separate releases on several new insecticides. The booklet described the diagnosis and treatment of persons who may have had extensive or intensive exposure to economic poisons. It was prepared primarily for the guidance of physicians and other public health professionals. The 1963 booklet concerned use of organophosphorous insecticides and acute toxicities associated with pesticides such as “arsenic, thallium, phosphorous, and kerosene” because they were “leading causes of deaths associated with pesticides.” Hayes acknowledged the great potential value of the materials used as pesticides and urged the careful collection of clinical data and related information concerning poisoning, a theme that became much clearer in the expanded Toxicology of Pesticides (1975). Toxicology of Pesticides and his works that followed gave attention to “those materials that are manufactured in

large amounts, that are known to have caused poisoning relatively frequently, or that are of special interest for some other reason.” The subjects of clinical studies included: (1) persons with “heavy occupational exposure”—including malaria control spray operators, farmers, orchardists, spray pilots, and pest control operators; (2) volunteers who take part in strictly controlled experimental investigations; and (3) patients who are sick from accidental over-exposure to pesticides. In the preface to his next major work and the first edition in the present series, he called attention to the need for basic toxicology education. Pesticides Studied in Man (1982) and The Handbook of Pesticide Toxicology represent his commitment to the collection and dissemination of critical research and clinical experience in Hayes’ career as a leader in pesticide science. Widespread use of the Clinical Handbook on Economic Poisons and active participation in public debate concerning pesticide use encouraged Hayes to write of the general importance of principles of toxicology. In Toxicology of Pesticides (1975) and his subsequent books he retained the strong clinical content but offered much expanded coverage of principles of toxicology, the conditions of exposure, the effects on human health, problems of diagnosis and treatment, the means to prevent injury, and even brief outlines on the impact of pesticides on domestic animals and wildlife. In the public arena, Hayes spoke out on an expanding role of toxicology to address issues of public and environmental health related to pesticide use that became critical during the 1960s and 1970s following publication of Rachel Carson’s polemic Silent Spring (1962). Concerning v

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Dedication

vi

the resulting intense public debate about pesticides, Hayes wrote in the Preface to Toxicology of Pesticides: “The pesticide problem is not merely one concerning the chemical industry and professional farmers, foresters, and applicators, or one concerning only those who wish to protect wildlife, or those responsible for control of malaria and other vector-borne diseases of man and his livestock. Rather, the pesticide problem concerns every person who wants food at a reasonable price and who wants his home free from vermin. The problem can be solved only on the basis of sound toxicological principles. Knowledge of these principles permits agreement and a cooperative approach on the part of persons professionally responsible for protection of our food, our health, and our wildlife, respectively. Ignorance of these principles limits some other persons to a partisan approach that may be dangerous to the common good.”

In dedicating Toxicology of Pesticides to Paracelsus, Hayes sought to bring attention to the “decisive importance of dosage” in determining the effect of exposure. He urged recognition of “tolerated doses” as well as information on doses or blood levels that have produced harm. He clearly viewed modern toxicology as a predictive, interdisciplinary science with great capacity to contribute to chemical safety evaluation. His Pesticides Studied in Man (1982) assumed the reader’s mastery of the basic principles of toxicology and offered more in-depth coverage of those pesticides with direct information concerning their effects in humans. The information came from reports of poisoning, from observation of workers or volunteers, or from persons who received certain compounds as drugs. Sections were organized in three parts. The first gave a concise summary of the chemistry and use of the pesticide. The second part concerned the fate and basic animal toxicity data that contributed to determining important dose-response relationships. The third section reported the human experience with the pesticide. The present edition of Hayes’ Handbook of Pesticide Toxicology applies this basic scheme more loosely in the description of the toxicology of agents. As Professor of Biochemistry, School of Medicine, Vanderbilt University, Hayes teamed with his colleague Edward R. Laws, Jr., Department of Neurological Surgery, George Washington School of Medicine, Washington, D. C. to edit the first edition of the Handbook of Pesticide Toxicology. It was published by Academic Press in three volumes and updated and revised both Toxicology of Pesticides and Pesticides Studied in Man. The Preface again champions the potential role of toxicology in resolution of controversy regarding pesticide use and reiterates the importance of the study of dose-response relationships in diagnosis of poisoning. The book follows familiar organization including exposition of principles of toxicology and sections featuring the chemistry and uses of pesticides,

biochemistry and experimental toxicology, and description of the human experience with pesticides. Hayes’ admonition to physicians to collect quantitative information on the effects of different dosages is consistent with his high regard for the fullest possible data concerning the human experience with pesticides. Throughout his career Hayes shaped a vision of modern toxicology as an important means to achieve rational use of chemicals in the environment, much in the spirit of Paracelsus who wrote, “… whenever I went I eagerly and diligently investigated and sought after the tested and reliable arts of medicine. I went not only to the doctors, but also to barbers, bathkeepers, learned physicians, women, and magicians who pursue the art of healing.” Wayland Hayes was born in Charlottesville, Virginia, on April 29, 1917. He graduated in 1938 from the University of Virginia, received an M. A. degree and a Ph. D. from the University of Wisconsin where he specialized in zoology and physiological chemistry. He returned to the University of Virginia where he received the M. D. in 1946. He interned in the Public Health Service Hospital in Staten Island, New York, and entered the regular corps of the service from 1948 to 1968. He became Chief Toxicologist of the Pesticides Program of the Centers for Disease Control in Savannah and Atlanta, Georgia. Hayes joined Vanderbilt University as Professor of Biochemistry, School of Medicine, in 1968 becoming emeritus in 1982 but remaining active in university affairs until 1991. He died January 4, 1993. His wife, Barnita Donkle Hayes, of 50 years and a son, Wayland J. Hayes III; and four daughters, Marie Royce Hayes, Maryetta Hayes Hacskaylo, Lula Turner McCoy and Roche Del Moser; and 10 grandchildren, survived him. In his family and community, he was revered as a parent, gardener, artist, philosopher and humorist. Hayes had a full professional life of national and international service. He was a consultant on the toxicology of pesticides to the World Health Organization, the Pan American Sanitary Bureau, the American Medical Association, the U. S. Department of Agriculture/Environmental Protection Agency, the American Conference of Governmental Industrial Hygienists and the National Academy of SciencesNational Research Council. He served on numerous governmental committees and editorial boards. He was a charter member of the Society of Toxicology in 1961 and served as its eleventh president 1971–72. As president of the Society, he staunchly defended the integrity of toxicologists in regulatory affairs (Science 174: 545–546, 1971) and launched criticism of the USEPA’s dismissal of the recommendation of its own Scientific Advisory Committee in response to “external pressure.” As president, Hayes made a strong plea for the inclusion of toxicology in textbooks of biology, zoology, hygiene, and general science (Toxicology and Applied Pharmacology 19, i–ii, 1971). Both subjects are topical today. Other society memberships included the American

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Dedication

vii

Society of Pharmacology and Experimental Therapeutics and the American Society of Tropical Medicine and Hygiene. He became a Diplomat of The Academy of Toxicological Sciences in 1989. Wayland Hayes was a sought after expert witness particularly in cases involving pesticides. His commanding and distinguished presence, his southern accent and gracious manner coupled with his encyclopedic knowledge rarely failed to win the case. However, there was one case in Wisconsin where he was unable to convince the jury that DDT was not a potent poison. Finally, he walked over to the evidence table, picked up the bottle of DDT and ingested a teaspoon of the evidence. When asked about how that worked out, he replied, “well I may have walked a little funny, but we won the case”.

Hayes clearly recognized the difficulties associated with collecting meaningful dosage-response information. He suggested that failure to collect such valuable data might result from lack of recognition of its importance in diagnostics. He closed on a theme that has shaped his career and that remains central to the spirit and content of the current volumes now dedicated to his life and career saying, “Clinicians who attend patients poisoned by a pesticide or by any other material are urged to be alert to the possibility of getting new information on dosage.”

(c) 2011 Elsevier Inc. All Rights Reserved.

Robert I. Krieger, Ph.D. John W. Doull, M.D., Ph.D.

Contents of Volume 1 Contributors Foreword Preface

xxiii xxix xxxvii

Section I Pesticide Uses 1.

2.

Dose and Time Determining, and Other Factors Influencing, Toxicity

3

1.1

3

Introduction 1.1.1 1.1.2 1.1.3 1.1.4

1.2

1.3

10 10 10

Quantitation of dosage–response relationships

1.3.5 1.3.6 1.3.7

1.5

Nature of the Injury Duration of the Injury

ED 50 or LD 50 Measurement of Cumulative Effects Time Relationships Problem of Measuring Effect of Dispersed Toxicants Measurement of Graded Responses Dosage at the Tissue Level Statistical Considerations

2.8

2.9 2.10 2.11

23 27

2.12 2.13 2.14

33 33 33 34

47

1.4.1 1.4.2 1.4.3 1.4.4 1.4.5 1.4.6 1.4.7 1.4.8

47 47 47 48 50 50 50 51

3.

3.1 3.2

3.3

3.4

Dosage Compound Interaction of Compounds Schedule of Dosage Duration of Dosage Route of Exposure

53 54 56 61 62 63

3.5

64 88

103 103 103 105 105 105 106 107 108

Herbicides (Table 2.3, Figure 2.4) 108 Fungicides and Insecticides (Table 2.4, Figure 2.5) 109

Respiration (Table 2.5, Figure 2.6) Growth regulators (Table 2.6, Figure 2.7) Unknown, nonspecific and other targets (Table 2.7) Overview (Table 2.8) Conclusion Postscript Acknowledgments

Pest Control Agents from Natural Products

Factors influencing toxicity of any kind 53 1.5.1 1.5.2 1.5.3 1.5.4 1.5.5 1.5.6

Introduction Primary targets Secondary targets Common target for structurally diverse pesticides Resistance as a limiting factor Nerve (Table 2.1, Figure 2.2) Photosynthesis and pigment synthesis (Table 2.2, Figure 2.3) Biosynthesis 2.8.1 2.8.2

13

Dosage–response relationships in different kinds of toxicity or change Toxicity (Sensu Stricto) Neurotoxicity Teratogenesis Carcinogenesis Mutagenesis Hypersensitivity and Allergy Induction of Enzymes Metabolism and Storage

2.5 2.6 2.7

12

Species and Strain Differences Discussion of Factors Influencing Toxicity

Pest Toxicology: The Primary Mechanisms of Pesticide Action 2.1 2.2 2.3 2.4

3 6 6 8

Kinds of toxicity

1.3.3 1.3.4

1.4

Dose and Time as Fundamental Variables of Toxicity Definition of Dose and Time Dose and Time Relationships Analogy to Thermodynamics

1.2.1 1.2.2

1.3.1 1.3.2

1.5.7 1.5.8

111 112 112 113 116 116 116

119

Introduction Insect control agents

119 120

3.2.1 3.2.2 3.2.3

120 154 167

Botanical Insecticides Microbial Insecticides Semiochemicals

Disease control agents

173

3.3.1 3.3.2

173 180

Fungicides Bactericides

Herbicides

183

3.4.1 3.4.2

183 184

Bilanafos (Bialaphos) Glufosinate

Rodenticides

188

3.5.1 3.5.2 3.5.3 3.5.4

188 193 196 200

Strychnine Red Squill and Scilliroside Ricin Salmonella Bacteria

ix

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x

4.

Public Health Pesticides 4.1 4.2 4.3 4.4

231

Introduction Definition of terms in vector-borne diseases Impact of arthropods on human health Integrated pest management and vector management 4.4.1 4.4.2

Noninsecticidal Methods in Vector Management Chemicals in Vector Management

Conclusion

5.

231 231

Pesticide Use and Associated Morbidity and Mortality in Veterinary Medicine 7.1 7.2 7.3 7.4 7.5

237 238 240

7.6 7.7 7.8

241

250

The Changing Role of Insecticides in Structural Pest Control 257 5.1 5.2 5.3 5.4

5.5 5.6 5.7

6.

7.

Introduction Pest problems: real or perceived Environmental and health concerns Insecticide applications

257 257 258 260

5.4.1 5.4.2

260 261

Nonresidual Insecticides Residual Insecticides

Soil treatments for subterranean termites Baits and baiting Future directions Conclusion

263 265 266 266

6.6 6.7 6.8 6.9 6.10

Introduction Vertebrate pests: what are they? Management restrictions Problems created by vertebrates Nonlethal management without pesticides Population reduction without pesticides Pesticides: repellents versus lethal agents Repellents Immobilizing agents Lethal vertebrate pesticides

274 275 277 277

6.10.1 6.10.2 6.10.3 6.10.4 6.10.5 6.10.6 6.10.7 6.10.8 6.10.9

277 281 281 282 282 282 283 283 283

Poison Rodent Baits Fumigants Tracking Powders Contraceptives Glue Boards Livestock Protection Collars Toxicant Ejector Device Flock Dispersal Agent Poison Bird Bait

Conclusion

285 286 286 287

7.8.1 7.8.2 7.8.3

290 290

Acute Intoxication Chronic Intoxication Pesticide Use and Cancer in Animals Pesticide Use and Exposure of People in Contact with Animals

Major pesticide categories

287 289 289 290

291 291

292

7.9.1

Vertebrate Pest Control Chemicals and Their Use in Urban and Rural Environments 271 6.1 6.2 6.3 6.4 6.5

Introduction Formulations Species sensitivities Pesticide use in domestic animals Regulation of pesticides used in veterinary medicine Violative residues Frequency of intoxication Scenarios of concern

7.8.4

7.9

285

271 271 272 272

7.10 7.11

Cholinesterase Inhibitors: Organophosphoruses and Carbamates 7.9.2 Pyrethrins and Pyrethroids 7.9.3 Natural Products Used for Flea Control 7.9.4 Macrocyclic Lactones 7.9.5 Neonicotinoids: Imidacloprid and Nitenpyram 7.9.6 Fipronil 7.9.7 Amitraz 7.9.8 Insect Growth Regulators 7.9.9 Synergists and Repellants 7.9.10 Rodenticides 7.9.11 Metaldehyde 7.9.12 Paraquat

294 295 295 295 295 296 296 296

Diagnosis of intoxication Treatment of intoxication Conclusion

296 297 298

292 292 293 294

272 273

284

8.

Pesticide Use Practices in Integrated Pest Management 8.1 8.2 8.3 8.4 8.5

8.6

Integrated pest management What is integrated pest manangement? The IPM continuum Pesticides Field scouting

304

8.5.1 8.5.2

Monitoring Decision Support

306 307

Reduced-risk pesticides

307

8.6.1 8.6.2

8.7 8.8 8.9

303

Behavioral Chemicals 307 Conventional Products and Risk 308

Cultural and physical suppression Prevention Avoidance

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305 305 306 306

308 308 308

Contents of Volume 1

8.10 8.11

9.

xi

Pesticides and biological controls Advisory services Conclusion

Properties of Soil Fumigants and Their Fate in the Environment 9.1 9.2

9.3

9.4 9.5

9.6

309 309 311

315

9.2.1 9.2.2

318 319

319

9.3.1 9.3.2 9.3.3

319 320 320

Fumigant distribution in soil and efficacy assessment Currently available soil fumigants

321 322

9.5.1 9.5.2 9.5.3 9.5.4 9.5.5 9.5.6 9.5.7

322 323 323 324 324 325 325

Methyl Bromide Methyl Iodide (Iodomethane) 1,3-Dichloropropene Chloropicrin Methyl Isothiocyanate Dimethyl Disulfide Sodium Tetrathiocarbonate

Strategies to minimize emissions

326

9.6.1 9.6.2 9.6.3 9.6.4 9.6.5 9.6.6 9.6.7

326 326 326 326 327 327

Application Methods Plastic Films Irrigation or Water Treatment Chemical Amendment Organic Amendment Target Area Treatment Mass Balance of Fumigants Applied to Soil

Conclusion

327

11.3 11.4

11.5 11.6

12.

Toxicity and Safety Evaluation of Pesticides

12.1

12.2

12.3

333

Risk Assessment for Acute Exposure to Pesticides

337

10.1 10.2 10.3 10.4

Introduction Toxicological data Exposure data Examples

337 337 338 340

10.4.1 Methyl Parathion 10.4.2 Methyl Bromide

340 344

Conclusion

349

357

Introduction Genotoxicity tests

357 358

11.2.1 Bacterial Reverse Mutation Assay 11.2.2 In Vitro Mutation Assay in Mammalian Cells 11.2.3 In Vivo Cytogenetic Assay 11.2.4 Micronucleus Assay

360 361 361 362

Genotoxicity testing of pesticides Patterns of response

362 362

11.4.1 Pesticides Exhibiting Both Genotoxicity and Carcinogenicity 11.4.2 Pesticides Exhibiting Genotoxicity With Limited or No Evidence of Carcinogenicity 11.4.3 Pesticides Exhibiting Carcinogenicity Without Appreciable Genotoxicity 11.4.4 Nongenotoxic Agents Without Evidence of Carcinogenicity 11.4.5 Pesticides Exhibiting Mixed Results in Genotoxicity or Cancer Tests

367

Human biomonitoring Genotoxicity and risk assessment Conclusion

369 373 374

362

365

366 367

Developmental and Reproductive Toxicology of Pesticides 381

327

Section II Toxicity and Safety Evaluation

10.

11.1 11.2

317

Processes and factors affecting the fate of fumigants in soil Volatilization Degradation Adsorption

Genotoxicity of Pesticides

315

Introduction Chemical properties, application, and major environmental issues Fumigation Methods Environmental Concerns

11.

12.4

Introduction

381

12.1.1 Developmental Toxicity 12.1.2 Reproductive Toxicity 12.1.3 Epidemiology

382 383 383

Exposure

384

12.2.1 Timing of Exposure 12.2.2 Prenatal Reproductive Toxicants 12.2.3 Prepubertal Reproductive Toxicants 12.2.4 Adult Reproductive Toxicants

384 384 385 385

Mechanisms of action

385

12.3.1 Direct-Acting 12.3.2 Indirect-Acting

385 385

Regulatory issues

385

12.4.1 History 386 12.4.2 Principles of Testing and Evaluation 386 12.4.3 Choice of Species in Testing 388 12.4.4 Choice of Testing Doses 388 12.4.5 Interpreting Effects 388 12.4.6 Statistical Evaluation 389 12.4.7 Exposure Assessment 389 12.4.8 Impact of FQPA on Developmental and Reproductive Toxic Effects of Pesticides 390

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Contents of Volume 1

xii

12.5

Toxicology studies

391

12.5.1 12.5.2 12.5.3 12.5.4 12.5.5 12.5.6

391 399 414 414 421

Herbicides Insecticides Insect Growth Regulators Fungicides Rodenticides Animal Health Products, Fumigants, and Miscellaneous Pesticides

Conclusion

14.5

421

427

13.3

14.

Introduction Toxicity testing requirements for MPCAs

441

13.2.1 Tier I 13.2.2 Tier II 13.2.3 Tier III

445 446 446

Toxicity of individual MPCAs

447 447 454 456

Conclusions Acknowledgments

457 458

The Assessment of the Chronic Toxicity and Carcinogenicity of Pesticides 14.1 14.2

14.3

463

14.2.1 14.2.2 14.2.3 14.2.4

464 466 466 467

Assessment of chronic toxicity (noncancer) endpoints 14.3.1 Mortality 14.3.2 Clinical Observations 14.3.3 Body Weight, Body Weight Gain, and Food Consumption 14.3.4 Ophthalmoscopic Examination 14.3.5 Clinical Pathology 14.3.6 Organ Weight 14.3.7 Macroscopic and Microscopic Pathology 14.3.8 Additional Endpoints 14.3.9 Overall Assessment

14.4

15.3

15.4

15.5

15.6

15.7

467 468 468

472 473 473

Assessment of carcinogenic potential 474 14.4.1 Overall Approach 14.4.2 Statistical Considerations

15.2

464

468 469 469 472

474 475

477

15.8

16.

Challenges

479

14.6.1 Life Stages 14.6.2 Cumulative Toxicity

479 480

Conclusion

480

Immunotoxicity of Pesticides 15.1

463

Introduction Regulatory requirements, test guidelines, and protocols Species and Study Duration Route of Administration Dose Levels Chemical Purity

15.

445

13.3.1 Bacteria 13.3.2 Eukaryotic Pesticides (Fungi and Nonfungi (Stramenopila)) 13.3.3 Viruses

Application to risk assessment and regulatory decision making

Chronic Toxicity Risk Assessment – Noncancer Endpoints 478 14.5.2 Cancer Risk Assessment – Threshold Carcinogens 478 14.5.3 Cancer Risk Assessment – Nonthreshold Carcinogens 479

Microbial Pest Control Agents: Use Patterns, Registration Requirements, and Mammalian Toxicity 441 13.1 13.2

475 477

14.5.1

14.6

13.

14.4.3 Mode of Action 14.4.4 Human Relevance

483

Introduction

483

15.1.1 Immune System 15.1.2 Pesticides

483 484

Carbamates

484

15.2.1 Rodent Studies 15.2.2 Nonrodent Studies 15.2.3 Human Studies

485 486 486

Organochlorines

486

15.3.1 Rodent Studies 15.3.2 Nonrodent Studies 15.3.3 Human Studies

486 487 488

Organophosphates

488

15.4.1 Rodent Studies 15.4.2 Nonrodent Studies 15.4.3 Human Studies

488 489 489

Phenoxy Compounds

489

15.5.1 Rodent Studies 15.5.2 Human Studies

490 490

Pyrethroids and Pyrethrins

490

15.6.1 Rodent Studies 15.6.2 Nonrodent Studies 15.6.3 Human Studies

490 490 491

Triazines

491

15.7.1 Rodent Studies 15.7.2 Nonrodent Studies 15.7.3 Human Studies

491 491 492

Regulations

492

15.8.1 Immuntoxic Guidelines 15.8.2 Immunotoxic Testing

492 492

Conclusion Acknowledgments

493 493

Risk Assessment for Acute, Subchronic and Chronic Exposure to Pesticides: Endosulfan 499 16.1 16.2 16.3 16.4

Introduction Chemical identification Environmental fate Mechanism of toxicity

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499 500 500 501

Contents of Volume 1

16.5 16.6 16.7

16.8

16.9

xiii

Biotransformation Pharmacokinetics Toxicology profile

501 501 501

16.7.1 Acute Toxicity 16.7.2 Subchronic Toxicity 16.7.3 Chronic Toxicity and Oncogenicity 16.7.4 Genotoxicity 16.7.5 Reproductive Toxicity 16.7.6 Developmental Toxicity 16.7.7 Neurotoxicity

501 502

17.2

17.2.1 Pesticides, Genes, and Cancer Risk 17.2.2 Pesticides, Genes, and Parkinson’s Disease

17.3

503 503 503 504 504

Hazard Identification

505

16.8.1 16.8.2 16.8.3 16.8.4

505 506 506 506

17.4

Exposure Assessment

506

16.9.1 Occupational 16.9.2 Bystanders at Application Sites

506 510

Conclusions

510

16.10.1 DPR Residue Database and Exposure Analysis 16.10.2 Tolerances in California 16.10.3 Residue Adjustments: Percentage of the Crop Treated 16.10.4 Highest Measured Acute Residue Values

16.11 Occupational and bystander air aggregate exposure

18.

510 512 512 512

513

16.11.1 Aggregate (Dermal  Inhalation  Dietary) Exposure in Occupational Scenarios 513

16.12 Risk characterization

16.13 Issues related to sensitive populations 16.14 Endocrine disruption and the FQPA safety factor 16.15 Tolerance assessment updated by DPR for 2006 16.16 Summary

515 516 516

19. 516 516

Conclusion

516

Introduction

18.5

514

16.16.1 Occupational and Bystander MOEs 16.16.2 Dietary MOEs

Genetic Polymorphism and Susceptibility to Pesticides

18.4

514 514

525 525

525 526

529 536 538 539 540

541

551

Introduction Steroid hormones

551 553

18.2.1 18.2.2 18.2.3 18.2.4 18.2.5

553 554 555 557 557

Estrogen Receptor Function Androgen Receptor Function Steroidogenesis Inhibitors Inducers of Steroid Clearance Enhancers of Steroid Action

Hypothalamic–Pituitary–Gonadal Axis

558

18.3.1 Dithiocarbamates 18.3.2 Atrazine 18.3.3 Pyrethroids

558 558 559

Thyroid Hormone

560

18.4.1 2,4-D 18.4.2 Triclosan

560 561

Impact on Testing Guidelines

561

18.5.1 Multigenerational Studies 18.5.2 Endocrine Disruptors Screening Program

561 562

Conclusion Acknowledgments

565 565

Volatile Organic Compounds from Pesticide Application and Contribution to Tropospheric Ozone 19.1 19.2 19.3

Section III Emergining Approaches in Safety Evaluation

17.1

18.3

513

16.12.1 MOE Calculations 16.12.2 MOE Results

Cytochromes P450S Paraoxonase Additional Carboxylic Esterases Glutathione S-Transferase Additional Putative Pesticide Susceptibility Genes

Pesticides as EndocrineDisrupting Chemicals 18.1 18.2

525

Importance of environmental exposure assessment in g  e studies of pesticides 528 Specific genes and polymorphisms relevant to putative gene–pesticide interactions 529 17.4.1 17.4.2 17.4.3 17.4.4 17.4.5

Acute Toxicity Subchronic Toxicity Chronic Toxicity Genotoxicity/Oncogenicity

16.10 Dietary Exposure

17.

Diseases with putative “gene–pesticide” interactions

571

Introduction Principal voc contributing pesticides CDPR initial method to obtain a screening assessment of voc mass released from each california pesticide—early 1990s

571 572

19.3.1 Method Development 19.3.2 Method Validation 19.3.3 Corrections to TGA Results

572 572 573

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572

Contents of Volume 1

xiv

19.4

19.3.4 Default Volatility Values 19.3.5 VOC Inventory Calculations

573 573

Revisions and Improvements to the Early Program

574

21.2.2 Chromatographic Separation of Proteins 21.2.3 Stable Isotope Labeling of Proteins 21.2.4 Mass Spectrometry Identification of Proteins 21.2.5 Protein Bioinformatics

19.4.1 Special Default Emission Potential 574

19.5

19.6

20.

Fumigant Adjustments for Field Application Method

574

19.5.1 19.5.2 19.5.3 19.5.4 19.5.5 19.5.6

574 574 575 575 575 575

Methyl Bromide (MB) 1,3-Dichloropropene (1,3-D) Chloropicrin (CP) Metam Sodium Dazomet Sodium Tetrathiocarbonate

Future Issues

575

19.6.1 Adjusting Nonfumigants for Application Method Effects on Emissions 19.6.2 Speciation/Reactivity 19.6.3 Ozone Toxicity 19.6.4 Mechanisms of Toxicity 19.6.5 Animal Studies 19.6.6 Human Studies

575 576 577 577 577 579

Conclusion

581

Regulatory Aspects of Acute Neurotoxicity Assessment 20.1 20.2 20.3

20.4 20.5

587

20.3.1 20.3.2 20.3.3 20.3.4

587 589 590 591

Standard Acute Toxicity Studies Regulatory Neurotoxicity Studies 20.5.1 Delayed Neurotoxicity of Organophosphorus Compounds 20.5.2 Neurotoxicity Screening Battery/ Neurotoxicity Study 20.5.3 Schedule-Controlled Operant Behavior 20.5.4 Peripheral Nerve Function 20.5.5 Neurophysiology: Sensory Evoked Potentials 20.5.6 Acute Cholinesterase Comparative Sensitivity Studies

Conclusion

21.

21.4

Conclusion

22.

22.1 22.2

22.3

607 607

609 611 612 618 618 619

619

Introduction Metabolism, Metabolites, and Metabolomics

627

22.2.1 Metabolism and Metabolites 22.2.2 Metabolomics

627 628

Research Methods in Metabolomics

629

22.3.1 Sample Pretreatment and Extraction Methods 22.3.2 Qualitative and Quantitative Measurements 22.3.3 Data Mining Methodologies 22.3.4 Emerging Technologies With Novel Applications

587 587

22.4

591 592

606

Metabolomics in Pesticide Toxicology 627

Applications Conclusions Acknowledgments

627

630 631 634 635

636 638 639

592 594 596 597

Section IV Dermatotoxicology of Pesticides 23.

598 599

Irritant Dermatitis 23.1 23.2

599

Proteomics in Pesticide Toxicology 603 21.1 21.2

Applications of Proteomics in Pesticide Studies Microbial Degradation of Pesticides 21.4.1 Proteomics in Bacterial Degradation of Pesticides 21.4.2 Network of Catabolism and Central Metabolism 21.4.3 Bacterial Cell Membrane Proteins 21.4.4 Bacterial Stress Responses and Adaption

587

Introduction Toxicological Effects Following Acute Exposures to Pesticides Methodology for Assessing Neurotoxicity Behavioral Methods Electrophysiological Techniques Neurochemical Endpoints Neuropathological Methods

21.3

605

Introduction Proteomics Methods

603 603

21.2.1 Two-Dimensional Gel Electrophoresis

604

23.3

647

Introduction Factors Influencing Irritant Potential

647 647

23.2.1 Chemical Factors 23.2.2 Physical Factors 23.2.3 Endogenous Patient Characteristics

647 647 648

Identifying Suspected Irritants

648

23.3.1 Irritant Patch Testing 23.3.2 Cumulative Irritation Testing 23.3.3 Chamber Scarification Test

649 649 649

(c) 2011 Elsevier Inc. All Rights Reserved.

Contents of Volume 1

23.4

xv

23.3.4 Immersion Tests 23.3.5 Bioengineering Approaches 23.3.6 New Approaches

649 649 650

Clinical Patterns of Irritant Contact Dermatitis

650

23.4.1 Acute Irritant Dermatitis (Primary Irritation) 23.4.2 Delayed, Acute Irritant Dermatitis 23.4.3 Irritant Reaction 23.4.4 Subjective/Sensory Irritation 23.4.5 Suberythematous Irritation 23.4.6 Cumulative Irritant Dermatitis 23.4.7 Traumatic Irritant Dermatitis 23.4.8 Acneiform and Pustular Irritant Dermatitis 23.4.9 Exsiccation Eczematoid 23.4.10 Friction Dermatitis 23.4.11 Airborne Irritant Dermatitis

23.5 23.6 23.7

24.

24.1 24.2 24.3 24.4 24.5 24.6 24.7 24.8 24.9

25.

25.1 25.2 25.3 25.4 25.5 25.6

651 651 651 651

653 655

Conclusion

657

Allergic Contact Dermatitis Introduction Allergic Contact Dermatitis Prevalence of Allergic Contact Dermatitis Due to Agrochemicals Chemicals and Cases Diagnosis and Treatment of Allergic Contact Dermatitis Prevention Conclusion

26.1 26.2 26.3

650 650 650 651 651 651

23.7.1 Acute Agricultural Irritant Dermatitis 23.7.2 Acute Irritation from Pesticides (Fumigants and Insecticides) 23.7.3 Cumulative Irritant Dermatitis from Pesticides 23.7.4 Plants as Agricultural Irritants

Introduction Solar Radiation and Photosensitivity Photoirritant Reactions (Phototoxicity) Photoallergic Contact Dermatitis Pesticides That Cause Phototoxic or Photoallergic Reactions Approach to the Photosensitive Patient Phototesting Histology Management Strategies Conclusion

Pesticides and Contact Urticaria Syndrome

650

Diagnosis of Irritant Contact Dermatitis 651 Treatment of Irritant Contact Dermatitis 652 ICD in Agricultural Workers 653

Photocontact Dermatitis

26.

26.4 26.5 26.6

27.

27.1 27.2

27.3

661 27.4 27.5

664 665 666 666 667

669 669 669

672 673 674

Animal and Human Assays Diagnosis of CUS CUS Induced by Pesticides Conclusion

679 679 679 680

Introduction Percutaneous Absorption Methodology

Regional Variation in Human and Animal Pesticide Percutaneous Absorption Percutaneous Absorption from Chemicals in Clothing Models for Agricultural Chemical Assessments and Predictions 27.5.1 The Cluster Analysis Method 27.5.2 Determinants of Dermal Exposure Ranking Method (DERM): A Method for Pesticide Exposure Assessment for Developing Countries 27.5.3 Dermal Assessment Estimate (DREAM) Method 27.5.4 Estimation and Assessment Exposure (EASE) Model 27.5.5 Risk Assessment of Occupational Dermal Exposure (RISKOFDERM) 27.5.6 Biosensors

664

670 670

678 678

27.2.1 Absolute Topical Bioavailabillity 27.2.2 Radioactivity in Excreta 27.2.3 Skin Flaps 27.2.4 Stripping Method 27.2.5 Biological Response 27.2.6 In Vitro and In Vivo Methodologies

653

27.6 27.7

677 677 678

26.3.1 Immunologic Contact Urticaria 26.3.2 Nonimmunologic Contact Urticaria 26.3.3 Uncertain Mechanism

Agricultural Chemical Percutaneous Absorption and Decontamination

653

661 661 661 662

Definition Clinical Signs and Symptoms Etiology and Mechanisms

677

Biomonitoring: Tool for Human Health Risk Characterization Skin Decontamination Conclusion

(c) 2011 Elsevier Inc. All Rights Reserved.

678

683 683 683 683 683 685 685 686 686

687 689 692 692

694 695 695

696 696

696 696 698

Contents of Volume 1

xvi

28.

The Regulatory Evaluation of the Skin Effects of Pesticides 28.1

28.2

28.3

32. 701

Introduction

701

28.1.1 Basic Patterns of Skin Reaction 28.1.2 Testing Requirements and Test Protocols 28.1.3 Regulatory Decisions 28.1.4 Integration of Illness Surveillance Data with Experimental Dermal Irritation and Sensitization Tests

701

704

Review of Use Categories

706

32.1 32.2

701 704

28.2.1 Antimicrobial Agents/ Disfinfectants 28.2.2 Insecticides and Insect Repellants 28.2.3 Fungicides 28.2.4 Fumigants and Biocides 28.2.5 Herbicides

735 753 764 767

Adjuvants Conclusion

781 781

32.3

706

Neurotoxicology of Pesticides

30.

A Systems Biology Approach to Assess the Impact of Pesticides on the Nervous System 30.1 30.2 30.3 30.4

31.

31.1

31.2

799

837 837 839 841 843 843

847

34.2.1 Types of Animal Models

848

34.3

Considerations for Animal Models of Pesticide Exposures

848

34.4

Overview and Summary Acknowledgments

808

35. 36.

837

847 847

Section VI Pesticides Disposition

809 809 811

830

Introduction Animal Model Considerations

806

808

822 826 828 828 828

34.1 34.2

34.3.1 Doses, Timing, and Duration of Exposures and Endpoint Considerations 34.3.2 The Young as Sensitve Targets 34.3.3 Endpoints of Toxicity 34.3.4 Nonhuman Primate Behavioral Concordance with Humans 34.3.5 Automated Assessments

805 806

Introduction Human effects Rat models Mouse models Fish models Conclusions

The Nonhuman Primate as a Translational Model for Pesticide Research

799 799

Fipronil Imidacloprid Acknowledgments

34.

793 794 794 794 796

31.1.1 Sodium Channel Modulation 31.1.2 Pyrethroid Action on Other Receptors and Channels 31.1.3 Sodium Channel Mutation in Pyrethroid Resistance

Cyclodienes and Hexachlorocyclohexane

Lasting Behavioral Consequences of Organophosphate Pesticide Exposure During Development 33.1 33.2 33.3 33.4 33.5

793

Pyrethroids and DDT

31.2.1 GABAA Receptor Subunit Specificity of Dieldrin Action 31.2.2 GABAA Receptor Subunit Specificity of The Actions of HCH Isomers

31.3 31.4

791

Introduction Pesticides and Developmental Toxicity Developmental Neurotoxicity Examples from the Current Literature Conclusions

Neurophysiological Effects of Insecticides

Organophophorus Pesticides Organochlorine Insecticides Carbamates Pyrethroid Insecticides Miscellaneous Pesticides

Conclusion

33.

819

General Concepts in Differential Sensitivity to Pesticides 819 Children’s Health and Regulation of Pesticides in the United States 821 Age-Related Differences in Sensitivity to Pesticides 822 32.3.1 32.3.2 32.3.3 32.3.4 32.3.5

Section V Neurotoxicology of Pesticides 29.

The Influence of Age on Pesticide Toxicity

Introduction to Pesticide Disposition

849 849 850 851 852

855 856

863

Introduction to Biotransformation (Metabolism) 865 36.1

Introduction

(c) 2011 Elsevier Inc. All Rights Reserved.

865

Contents of Volume 1

36.2

xvii

Xenobiotic-Metabolizing Enzymes 36.2.1 Cytochrome P450 Monooxygenases 36.2.2 Flavin-containing Monooxygenase 36.2.3 Other Phase I Reactions 36.2.4 Phase II Reactions: Conjugations

36.3

37.

Absorption 37.1 37.2

37.3

37.4

38.

38.4 38.5

38.6

866 867 869

39.3 871

873 873

877

Introduction Factors that Influence the Transfer and Availability of Chemicals in the Body

877

37.2.1 Properties of Cell Membranes 37.2.2 Transport Mechanisms 37.2.3 Protein (Macromolecular) Binding

878 878 879

Absorption

880

37.3.1 Percutaneous Absorption 37.3.2 Absorption From the Gastrointestinal Tract 37.3.3 Absorption From the Respiratory Tract 37.3.4 Absorption After Exposure by Other Routes

881

888

Summary and Future Directions

888

Metabolism of Pesticides 38.1 38.2 38.3

39.

Major xenobiotic Biotransformation Reactions Conclusions

40. 878

40.3

887

Introduction External Transformation Biotransformation

893 893 894

38.3.1 Biotransformation in the Liver 38.3.2 Biotransformation in Extrahepatic Tissues

898

Toxicity of Metabolites Physiological Factors Affecting Biotransformation

909

38.5.1 38.5.2 38.5.3 38.5.4 38.5.5

910 910 910 910 911

Developmental Effects Species Differences Individual and Strain Differences Gender Differences Genetic Factors

40.2

40.4

893

907

40.5 40.6

41.

41.1 41.2

Tolerance and Resistance

911 911 911

Conclusions

912

Introduction Distribution

923 923

39.2.1 Total Body Water

924

41.3 41.4 41.5

41.6 41.7

924 924 924 925 925 925

Pharmacokinetics

926

39.3.1 Noncompartmental Models 39.3.2 Overview of Classical Compartmental Models 39.3.3 One-Compartment Model 39.3.4 Multicompartment Models 39.3.5 Physiologically Based Models

926 928 928 932 934

Conclusion

938

941

Chemical Factors Affecting Pesticide Metabolism: Introduction Induction

941 941

40.2.1 Induction of Microsomal Enzyme Activity 40.2.2 Induction of Other Enzymes 40.2.3 Mechanism of Induction

941 948 948

Inhibition

948

40.3.1 Types of Inhibition and Experimental Demonstration 40.3.2 Synergism and Potentiation 40.3.3 Antagonism 40.3.4 Pesticides as Inhibitors

948 950 950 951

Biphasic Effects: Inhibition and Induction Activation Hepatotoxicity Conclusion

951 953 953 953

Pesticide Excretion

910

38.6.1 Tolerance 38.6.2 Resistance

Rate and Extent of Distribution Volume of Distribution Blood–Brain Barrier Placental Transfer Storage and Redistribution Storage with Repeated Exposure

Metabolic Interactions of Pesticides 40.1

886

Distribution and Pharmacokinetics 923 39.1 39.2

39.2.2 39.2.3 39.2.4 39.2.5 39.2.6 39.2.7

866

961

Introduction Renal function

961 961

41.2.1 41.2.2 41.2.3 41.2.4

961 962 962 962

Overall Aspects Glomerular Filtration Tubular Reabsorption Tubular Secretion

Biliary Excretion Respiratory Excretion Other Routes of Excretion

963 963 963

41.5.1 Gender-Linked Routes of Excretion 41.5.2 Alimentary Elimination 41.5.3 Obscure Routes of Excretion

963 964 964

Cellular Elimination Excretion of Pesticides and Their Metabolites as Biomarkers of Exposure Conclusion

(c) 2011 Elsevier Inc. All Rights Reserved.

964

965 966

Contents of Volume 1

xviii

Section VII Exposure Measurement and Mitigation 42.

Exposure Framework 42.1 42.2 42.3 42.4

43.

44.2

Sampling and Analysis for Nonoccupational Pesticide Exposure Assessments 43.1 43.2

43.3

43.4

44.

977

Introduction 977 Design of Nonoccupational Observational Exposure Measurment Studies 978 Sample Collection Methods 981

Analytical Methods for Pesticide Measurements

44.4

987

Conclusion Acknowledgments

991 992

989 991

Introduction

995

44.1.1 Terminology for Human Exposure Models

996

999 1002

Source-to-Dose Exposure Modeling for Pesticides 1002

Discussion: Models to Support Pesticide Regulation 44.4.1 Chromated Copper Arsenate (CCA) Exposures to Children 44.4.2 Assessing Chlorpyrifos Exposures and Doses

Conclusion

45.

Biomonitoring: Uses and Considerations for Assessing Nonoccupational Human Exposure to Pesticides 45.1 45.2 45.3

Introduction Linking Pesticide Exposure to Health Effects The Uses of Biomonitoring 45.3.1 Assessing Population-Based Exposure Trends 45.3.2 Improving Estimates of Exposure and Dose

45.4

987

43.4.1 General Principles 43.4.2 Method Performance Requirements 43.4.3 State-of-the-Science

998

44.3.1 Overview of the Steps in Exposure Modeling 44.3.2 Human Activity-Based Exposure Modeling 44.3.3 Toxicokinetic Modeling for Assessing Pesticide Uptake and Distribution 44.3.4 Exposure Reconstruction and Inverse Modeling 44.3.5 Sensitivity and Uncertainty Analyses

974 976

Modeling and Predicting Pesticide Exposures 995 44.1

44.3

971 971 973

43.3.1 Criteria for Selection of Sample Collection Methods 981 43.3.2 Methods for Estimating Inhalation Exposure 984 43.3.3 Methods for Estimating Dermal Exposure 984 43.3.4 Indirect Methods for Estimating Dermal Exposure 985 43.3.5 Methods for Estimating Indirect Ingestion 985 43.3.6 Direct Methods for Estimating Dietary Exposure: Duplicate Diet Samples 986 43.3.7 Water Collection Methods 986 43.3.8 Soil Measurement Methods 986 43.3.9 Collection of Biomonitoring Samples 986 43.3.10 Collection of Ancillary Information Such as Activity Data and Questionnaires 987

997

44.2.1 Screening-Level Exposure Models 44.2.2 Comprehensive Human Exposure Models

971

Introduction Exposure Science Exposure Assessment Considerations for Pesticide Exposure Assessments Conclusion

44.1.2 Pesticide Exposure Models

Types of Human Exposure Models

45.5

45.6

1003 1006

1008 1010 1011

1012 1013 1013

1016

1021 1021 1021 1022 1022 1026

Biomarker Selection and Use

1027

45.4.1 45.4.2 45.4.3 45.4.4 45.4.5

1027 1027 1029 1029 1029

Sensitivity Specificity Validity Biological Relevance Practicality

Factors Affecting the Use and Interpretation of Biomarkers of Exposure

1030

45.5.1 Kinetics 45.5.2 Urinary Excretion Rate

1030 1031

Summary and Suggestions for Future Research 1033 Conclusion 1034 Acknowledgments 1034

(c) 2011 Elsevier Inc. All Rights Reserved.

Contents of Volume 1

46.

Pesticide Exposure from Residential and Recreational Turf 1037 46.1 46.2 46.3 46.4 46.5 46.6

47.

Introduction Assessment of Transferable Turf Residues Activities on Residential and Recreational Turf Exposure Assessment Methodology Calculation of the Safe Residue Level Regulatory Approach to Postapplication Exposure to Turf Conclusion

47.2

47.3

48.1 48.2 48.3

1041 1042 1043 1043 1044

1047

47.1.1 Usage and Functional Aspects

1047

Turf Pesticides and Use

1048

47.2.1 Turf Pesticide Industry 47.2.2 Control of Turf Pests 47.2.3 Turf Pest Management

1048 1049 1055

Turf Pesticide Issues

1057

47.3.1 Environmental Issues 47.3.2 Exposure Issues 47.3.3 Best Management Practices

1057 1063 1069

Conclusions

1073

Introduction Pet Ownership and Pet Care in the United States Pet Care Products: How They are Used to Control Insect Pests 48.3.1 Exposure Monitoring Studies 48.3.2 Pet Care Product Applicator and Postapplicator Exposure Assessment Methods

48.4

49.

1039

Turf Environments

Pet Care Products Used for Insect Pest Control

Case study: Illustration of EPA Methodology

49.1

50.

50.1 50.2

50.3

1079

1084

Conclusion

1089

1087

1091 1091

Overview Exposure Models

Sample Model Calculation Methodologies 50.3.1 Typical Exposure Methodology Sequence 50.3.2 Monte Carlo Simulation Methodology 50.3.3 Degradation Calculation Methodology 50.3.4 Risk (MOE) Calculation Methodology

1082

1087 1087

1092 1093

1094 1094

1099 1101

50.2.1 General Exposure Model 1101 50.2.2 General Dietary Model 1102 50.2.3 Probabilistic or Monte Carlo Models 1103 50.2.4 Combination of Screening and Probabilistic Models 1103 50.2.5 Contact Factors for Dietary Exposure Model: Food Consumption 1103 50.2.6 Common Types of Food Consumption Surveys 1104 50.2.7 Estimating Residues in Foods 1106 50.2.8 Considerations for Cumulative Exposure Assessments 1106 50.2.9 Consideration for Aggregate Exposure Assessments 1107 50.2.10 Estimating Aggregate and Cumulative Exposures Using a Calendar Model 1107 50.2.11 Toxicity Data Used to Place the Exposure Assessment in Context 1108

1078 1079

Overview of General Issues Lessons Learned from Key Studies Guidance for Residential Postapplication Exposure Assessment Methods and Data Sources for Exposure Factors Research Needs

Modeling Dietary Exposure with Special Sections on Modeling Aggregate and Cumulative Exposure 1099

1077

1086

Introduction

49.5

1077

48.4.1 Properties of Formulation 48.4.2 Postapplication Dermal and Incidental Ingestion Exposure Variables 48.4.3 Activity Patterns 48.4.4 Research Data and Assessment Methodology Needs for Pet Care Pesticide Products

Residential Exposure Assessment: An Overview

49.2 49.3 49.4

1037

Lawn and Turf: Management and Environmental Issues of Turfgrass Pesticides 1047 47.1

48.

xix

50.4

50.5

1108 1108 1109 1109 1110

Uncertainty

1110

50.4.1 Scenario Uncertainty 50.4.2 Parameter Uncertainty 50.4.3 Model Uncertainty

1110 1110 1110

Example of Aggregate Exposure

1110

50.5.1 Description 50.5.2 Adult Applicator Dermal and Inhalation Exposures 50.5.3 Adult and Toddler Postapplication Dermal Exposures

1110

(c) 2011 Elsevier Inc. All Rights Reserved.

1111

1111

Contents of Volume 1

xx

50.5.4 Toddler Postapplication Incidental Ingestion Exposures 50.5.5 Reporting Results

50.6 50.7

51.

Quality Audit and Validation Commercial and Government Exposure Assessment Software

1112

50.7.1 Calendex 50.7.2 Cumulative and Aggregate Residue Evaluation System (CARES) 50.7.3 Dietary Exposure Evaluation Model (DEEM)-FCID 50.7.4 Dietary Exposure Potential Model (DEPM) 50.7.5 Lifeline 50.7.6 REx (Residential Exposure Assessment Model) 50.7.7 U.S. EPA Sheds-Wood Model

1112

1115 1115

Summary

1115

1112

52.7

1113 1113 1114 1114

53.

Introduction Method Validation

1117 1117

51.2.1 Confirmation of the Analytical Method 51.2.2 Independent Laboratory Validation 51.2.3 Method Radiovalidation (EPA, 1996)

51.3

1118 1118 1119 1120 1120 1123 1123

52.4 52.5 52.6

53.6 53.7

1127

Introduction Routes of Exposure

1127 1127

52.2.1 Dermal Exposure 52.2.2 Exposure by Inhalation 52.2.3 Oral Exposure

1127 1127 1128

Previous Reviews and Guidance on Methodology Design of Agricultural Worker Exposure Studies Test Subjects Methods for Measuring Exposure (Passive Dosimetry)

1128 1129 1130 1130

1135 1135

Conclusion

1136

1142

53.4.1 53.4.2 53.4.3 53.4.4 53.4.5

1142 1143 1143 1143 1146

Historical Background Overview of PHED PHED Grading Criteria Guidance for the use of PHED Limitations of PHED

1139 1139 1141

European Union Models and Databases

1146

53.5.1 Overview of the German and UK Operator Exposure Models (BBA Model and UK-POEM)

1146

The Agricultural Handlers Exposure Database (AHED) Re-entry Exposure Database

1149 1149

ARTF Cluster Survey Nine Special Clusters Eleven Field Crop Clusters Four Trellis Crop Clusters Four Orchard Crop Clusters Transfer Coefficient Statistical Model

Mitigation Measures for Exposure to Pesticides 54.1 54.2

1139

Introduction Basis for Generic Modeling The Tiered Approach The Pesticide Handlers Exposure Database (PHED)

Conclusion Acknowledgments

54.

1131 1131 1132 1134

52.7.1 Biological Monitoring

53.7.1 53.7.2 53.7.3 53.7.4 53.7.5 53.7.6

1124

Worker Exposure: Methods and Techniques

52.3

1118

51.3.1 51.3.2 51.3.3 51.3.4 51.3.5 51.3.6

Multiresidue Methods Extraction Cleanup of Extracts Separation and Quantitation Immunoassay Techniques Capillary Electrophoresis

53.5

1118

Development of the Analytical Method

Summary

52.1 52.2

1117

1130

Methods to Measure the Absorbed Dose

Operator and Field Worker Occupational Exposure Databases and Modeling 53.1 53.2 53.3 53.4

Modern Approaches to Analysis of Pesticide Residues in Foods and the Environment 1117 51.1 51.2

52.

52.6.1 Patch Method for Dermal Exposure 52.6.2 Use of Fluorescent Tracers and Visible Dyes: Quantification by Analysis or Video Imaging 52.6.3 Whole Body Method 52.6.4 Hand (and Head) Exposure 52.6.5 Inhalation Exposure

1112 1112

1151 1152 1152 1153 1153 1153

1153 1154

1157

Introduction Basics of a Pesticide EMM Process

1157 1158

54.2.1 Initiation of EMMs 54.2.2 Important Factors and Major Concerns

1159

(c) 2011 Elsevier Inc. All Rights Reserved.

1159

Contents of Volume 1

54.3

xxi

Criteria for More Practical, More Specific EMMs 54.3.1 EMMs for Pesticide Handlers/Users 54.3.2 EMMs for Fieldworkers 54.3.3 EMMs for Residents and Bystanders

54.4

Cases of Pesticide EMMs Actually Used 54.4.1 EMMs Actually Used by the U.S. EPA 54.4.2 EMMs Actually Used in California

Conclusion Acknowledgments

55.

Communicating Safe Pesticide Use 55.1 55.2

55.3 1162 55.4 1162 1163

55.5

1165

55.6

1166

55.7

1166

Who is the Appropriate Audience for Risk Communication? Perceptions of Risk: Why People Believe and Feel What They do Perceptions of Risk: Who Believes What and Whom? Needed Information: Some Ideas for Presenting Risk Concepts Benefits Information: The Most Needed Piece of the Communication Puzzle Conclusion

1166

1169 1170

1173

Why Pesticide Technology Users Need a Risk Communication Strategy 1173 Don’t Blame the Media 1174

(c) 2011 Elsevier Inc. All Rights Reserved.

1174 1176 1179 1181

1185 1186

Contents of Volume 2

Contributors

xxi

58.

Section VIII Regional and Global Environmental Exposure Assessments 56.

58.1 58.2

Ecotoxicological Risk Assessment of Pesticides in the Environment 56.1

Introduction

1191 1191

56.3

56.4

57.

The Risk Assessment Analysis

1199

56.2.1 Characterizing Effects 56.2.2 Characterizing Exposure

1200 1203

Risk Assessment of Pesticides

1206

56.3.1 Scoring Systems and Setting of Criteria 56.3.2 The Hazard Quotient 56.3.3 Probabilistic Risk Assessment

1206 1207 1207

Risk Communication Conclusion

1212 1213

Environmental Transport and Fate 57.1 57.2

57.3

1219

Introduction Principles

1219 1221

57.2.1 57.2.2 57.2.3 57.2.4

1221 1221 1222 1224

The Dissipation Process Environmental Compartments Structure Activation–Deactivation

Environmental Transport and Fate Modeling

1224

57.3.1 Physical Models 57.3.2 Mathematical or Computer Models

1224

Conclusion

1225

1225

1229

Introduction Measurements and Experimental Estimations of Log p

1229 1229

58.2.1 Direct Partitioning (Shake-Flask and Slow-Stirring) Method 1229 58.2.2 Potentiometric Titration Method for Ionizable Pesticides 1231 58.2.3 High-Performance Liquid Chromatographic Method 1231

56.1.1 Risk Assessment of Pesticides 1191 56.1.2 Assessing Risks from Pesticides in Relation to Other Substances 1193 56.1.3 Protection Goals, Assessment Endpoints, and Measures of Effects 1195

56.2

Hydrophobicity as a Key Physicochemical Parameter of Environmental Toxicology of Pesticides

58.3

Nonexperimental Estimations of Log P

1232

58.3.1 Scope and Limitation of the Additive Nature of Log P Values 58.3.2 Empirical (Manual) Procedure 58.3.3 Empirical Procedure Using Relationships with Log P Values of Simpler Compounds 58.3.4 Empirical Procedure Using Free-Energy Related Substituent Parameters 58.3.5 Computer-Aided Procedures

58.4

1232 1233

1234

1235 1237

Physicochemical Significance of Log P in Environmental Quantitative Structure-Activity Relationships 1239 58.4.1 Behavior in Soil 58.4.2 Bioaccumulation 58.4.3 Aquatic Toxicity

1239 1243 1245

Acknowledgments

1248

Section IX Public Health Regulation and Epidemiology 59.

Studies in Humans 59.1 59.2 59.3 59.4

1255

Cases Medical Use Use Experience Volunteers

1255 1256 1256 1256

59.4.1 Introduction 59.4.2 Legal and Ethical Considerations

1256 1258

v

(c) 2011 Elsevier Inc. All Rights Reserved.

Contents of Volume 2

vi

59.5

59.4.3 Design of Studies 59.4.4 Motivation of Volunteers 59.4.5 Studies of Pesticides in Volunteers 59.4.6 Conclusion

1271 1274

Measurement of Exposure and Dose Under Practical Conditions

1275

59.5.1 Measurement of Respiratory Exposure 59.5.2 Measurement of Dermal Exposure 59.5.3 Measurement of Oral Exposure 59.5.4 Problems of Measuring Separate Contributions from Different Routes of Exposure 59.5.5 Measurement of Absorbed Dose

59.6

Regulation of Pesticides and Other Chemicals by the EPA 59.6.1 EPA’s Risk Assessment Approach 59.6.2 Utility of Human Data in Risk Assessment 59.6.3 Attributes and Limitations of Laboratory Animal Data 59.6.4 Use of Toxicokinetic Data for Species-to-Species Extrapolations 59.6.5 Events Leading to NRC’s Review of the EPA’s Use of Data from Human Dosing Studies 59.6.6 EPA’s Charge to the NAS 59.6.7 NRC Committee and its Major Recommendations

Conclusion

60.

60.4

60.5

1275 1278 1280

1281

1299 1302

Chronic Poisoning by Pesticides

1302

60.5.1 Types of Chronic Pesticide Toxicity 60.5.2 Background Accumulation 60.5.3 Symptoms and Signs 60.5.4 Assessment of Exposure History 60.5.5 Assessment of Symptoms 60.5.6 Assessment of Signs 60.5.7 Workup of Neurotoxicity 60.5.8 Risks of Pesticide-Induced Cancer 60.5.9 Treatment

1308 1309

Conclusion

1309

1303 1304 1305 1305 1306 1306 1308

1285 1286 1286 1287

61.

Surveillance of PesticideRelated Illness and Injury in Humans 61.1 61.2

1288 1289 1289

1295

1296

60.3.1 60.3.2 60.3.3 60.3.4 60.3.5 60.3.6

1297 1297 1298 1298 1299 1299

1295 1295

Acute Poisoning by Pesticides

1299

60.4.1 Epidemiology

1299

1313

Introduction Surveillance Systems 61.2.1 American Association of Poison Control Centers’ National Poison data System 61.2.2 State-based Surveillance Systems 61.2.3 California Department of Pesticide Regulation 61.2.4 Bureau of Labor Statistics 61.2.5 Vital Status Statistics: Multiple Causes of Death 61.2.6 National Hospital Discharge Studies: Colorado State University 61.2.7 National Hospital Discharge Survey: National Center for Health Statistics 61.2.8 South Carolina Hospital Discharge Surveys 61.2.9 National Agricultural Workers Survey 61.2.10 Surveillance Efforts of International Organizations

1288

Introduction Types of Pesticides General Management of Acute Poisoning Skin Decontamination Activated Charcoal Gastric Lavage Cathartics Whole-bowel Irrigation Eye Contamination

60.4.2 Symptoms and Signs 60.4.3 Treatment

1282

1290

Diagnosis and Treatment of Poisoning Due to Pesticides 60.1 60.2 60.3

1264 1271

61.3

U.S. Environmental Protection Agency Regulations

1313 1314

1314 1325 1332 1339 1342

1343

1344 1345 1346 1347

1349

61.3.1 The Federal Insecticide, Fungicide, and Rodenticide Act 1349 61.3.2 Federal Reporting Requirements for Risk Information 1350

(c) 2011 Elsevier Inc. All Rights Reserved.

Contents of Volume 2

61.4 61.5

61.6

vii

61.3.3 National Pepticide Information Center 1350 61.3.4 Worker Protection Standard 1350

63.2

Evaluating Surveillance Systems Case Definition for Acute Pesticide-Related Illness and Injury

63.3

61.8

62.

1353

Limitations of Pesticide Poisoning Surveillance Data

1354

62.3

62.5

1356 1357

1360 1360

Internet and Telephone Resources for Pesticide Information Conclusion

64.

1363

64.1

65.

1364 1365

1371

65.3 65.4 65.5 65.6

Introduction 1371 Historical Background of Pesticide Regulation in the United States 1371 Current State of Pesticide Regulation in the United States 1372

1374

Current Regulatory Process

1377

62.4.1 Registration 62.4.2 Reregistration and Registration Renewal 62.4.3 Special Review

1377

Web Sites

1380

66.

Perceptions of Pesticides as Risks to Human Health

66.1 66.2

63.1

Introduction

1381

1395

64.1.1 History 64.1.2 Classification and Nomenclature 64.1.3 Synthesis 64.1.4 Reactions

1395 1395 1396 1398

1399

Introduction Oxidations

1399 1400

65.2.1 Cytochromes p450 65.2.2 Flavin Monooxygenases

1400 1401

Reductions Hydrolysis Conjugations Summary

1402 1402 1404 1404

1409

Background Pharmacokinetic Principles of Importance to Organophosphorus Insecticides 66.2.1 Compartmental Pharmacokinetic Models 66.2.2 Physiologically Based Pharmacokinetic Models

1378 1379

1381

Introduction

Organophosphorus Insecticide Pharmacokinetics

66.3

63.

1388 1390

The Metabolism of Organophosphorus Insecticides 65.1 65.2

1373

1384

Intuitive Toxicology: Expert and Lay Judgments of Chemical Risks Conclusion

Chemistry of Organophosphorus Insecticides 1395

1358

61.7.1 Principles of Epidemiology 61.7.2 Epidemiologic Study Designs 61.7.3 Evaluating Pesticide Health Information

1381

63.2.1 Social, Cultural, and Political Influences on Risk Perception

Section X Organophosphorous and N-Methyl Carbamate Insecticides

1357

1360

62.3.1 FIFRA—Key Changes and Additions 62.3.2 FFDCA—Key Changes and Additions

62.4

1354

Fundamentals of Epidemiology

Risk Assessment and Risk Management: The Regulatory Process 62.1 62.2

1352

61.5.1 The National Public Health Surveillance System Case Definition

61.6.1 Denominators 61.6.2 Limitations Related to Definitions 61.6.3 Limitations Related to Sensitivity 61.6.4 Legitimate uses for Surveillance Data 61.6.5 Mechanisms to Strengthen the Surveillance of Acute PesticideRelated Illness

61.7

1351

Risk-Perception Studies

Pharmacokinetic Approaches Applied to Organophosphorus Insecticides 66.3.1 Application of Pharmacokinetics to Understand the Overall

(c) 2011 Elsevier Inc. All Rights Reserved.

1409

1411 1411 1412

1415

Contents of Volume 2

viii

Disposition and Clearance of Organophosphorus Insecticides 1415 66.3.2 Development of Pharmacokinetic Models for Quantitative Biological Monitoring to Assess Organophosphorus Insecticide Exposure in Humans 1417 66.3.3 The Application of Pharmacokinetics for Quantifying Exposure to Organophosphorus Insecticides 1421 66.3.4 Studies that Facilitate Extrapolation of Dosimetry and Biological Response from Animals to Humans and the Assessment of Human Health Risk 1423

Conclusion Acknowledgments

67.4.3 Isolation and Localization 67.4.4 Molecular Biology 67.4.5 Structure of the Catalytic Domain

67.5

67.6

68.

68.4

67.

Neuropathy Target Esterase 67.1 67.2

Introduction Identication of Neuropathy Target Esterase

68.5 68.6 68.7

1435 1435

67.2.1 Organophosphorus Ester Insecticides: General Reactions with Serine Esterases 1435 67.2.2 Neuropathy Target Esterase as the Target for Initiation of Organophosphorus EsterInduced Delayed Neuropathy 1437 67.2.3 Possible Involvement of other Esterases in Organophosphorus Ester-Induced Delayed Neuropathy 1438

67.3

67.4

Toxicological Applications

1439

67.3.1 Hen Test 67.3.2 Structure–Activity Relationships and Prediction of Organophosphorus EsterInduced Delayed Neuropathy Potential in Hens 67.3.3 Application of Neuropathy Target Esterase Studies to Human Risk Assessment 67.3.4 Biomarkers and Biosensors

1439

1439

1440 1442

Nature and Properties of Neuropathy Target Esterase 1443 67.4.1 Biochemical Studies 67.4.2 Enzymology

1443 1443

Introduction Distribution Substrate Preferences and Selective Inhibitors Multiple Molecular Forms and Life History Mechanism of Hydrolysis Toxicities of Anticholinesterases Assay Techniques 68.7.1 Radiometric 68.7.2 pH 68.7.3 Thiol Substrates and the Ellman Assay 68.7.4 Variability

1435

68.8 Standards 68.9 Field Kits 68.10 Regulatory Matters: Are ChE Inhibitions Adverse Effects? 68.11 Blood ChEs and Detection of Exposure 68.12 Reactivation of Inhibited AChE 68.13 Significance of Blood ChEs 68.14 Direct Effects 68.15 Antidotes 68.16 Risk Assessment and ChEs Conclusion Acknowledgment

69.

Organophosphorus-Induced Delayed Neuropathy 69.1 69.2 69.3

69.4

1445

Role of Neuropathy Target Esterase in Organophosphorus Ester-Induced Delayed Neuropathy: A Toxic Gain of Function? 1447 Role of Neuropathy Target Esterase in Organophosphorus Ester-Induced Delayed Neuropathy: A Relative Loss of Function in Phospholipid Metabolism? 1448 Conclusion 1451

Cholinesterases 68.1 68.2 68.3

1429 1430

1444 1445

1457 1457 1457 1459 1459 1461 1463 1465 1465 1465 1466 1468

1469 1469 1471 1471 1471 1471 1472 1472 1473 1473 1474

1479

History Chemistry of Organophosphorus Compounds Clinical Manifestations

1480 1481

69.3.1 Human 69.3.2 Animal

1481 1481

Neuropathology

1482

(c) 2011 Elsevier Inc. All Rights Reserved.

1479

Contents of Volume 2

69.5 69.6 69.7

69.8

ix

Neuropathology of Mammalian Animal Models Pathogenesis Factors Influencing the Development of Organophosphorus-Induced Delayed Neuropathy Testing for OrganophosphorusInduced Delayed Neuropathy Conclusion

71.7.7 Reproductive and Developmental Effects 71.7.8 Neurotoxicity 71.7.9 Human Volunteer Study 71.7.10 Toxicology Studies on Malaoxon

1486 1490

1493

Chlorpyrifos 70.1 70.2 70.3

70.4 70.5

70.6

71.

1505

Introduction and Regulatory Aspects Physicochemical Properties Toxicokinetic Properties

1505 1505 1506

70.3.1 70.3.2 70.3.3 70.3.4

1506 1507 1507 1510

Absorption Distribution Metabolism Excretion

Exposure Toxicological Profile

1510 1511

70.5.1 70.5.2 70.5.3 70.5.4

Mechanism of Action Acute and Subchronic Toxicity Chronic Toxic Effects Genotoxicity and Carcinogeneticity 70.5.5 Reproductive and Developmental Toxicity

1511 1512 1514

Toxic Interactions Conclusion

1518 1519

71.1 71.2 71.3 71.4 71.5 71.6 71.7

71.7.1 Absorption, Distribution, Metabolism, and Excretion 71.7.2 Review of Toxicology Studies 71.7.3 Acute Toxicity 71.7.4 Short-Term and Subchronic Toxicity 71.7.5 Genotoxicity 71.7.6 Long-Term Studies

Clinical Toxicology of Anticholinesterase Agents in Humans

1543

72.2

72.3

72.4 1515

1527 1528 1528 1528 1529 1531

73.1

73.2

The Intermediate Syndrome

1564

72.2.1 Etiology 72.2.2 Pathogenesis 72.2.3 Clinical Manifestations

1564 1564 1565

Delayed Polyneuropathy

1565

72.3.1 Etiology 72.3.2 Pathogenesis 72.3.3 Clinical Manifestations

1566 1567 1567

Long-Term Exposures

1569 1569 1574 1575

Need for Cumulative Risk Assessments for n-Methyl Carbamate Insecticides Methodologies for Performing Cumulative Risk Assessments for N-Methyl Carbamates

1591

1593

73.2.1 Relative Potency Factor Approach Using an Index Chemical 1593 73.2.2 Advances in the PBPK/PD Modeling Approach 1594

1531

1533 1534 1535

1543 1543 1545 1547

Application of Physiologically Based Pharmacokinetic/Pharmacodynamic Modeling in Cumulative Risk Assessment for N-Methyl Carbamate Insecticides 1591

1531

1532 1532

The Cholinergic Syndrome 72.1.1 Etiology 72.1.2 Pathogenesis 72.1.3 Clinical Manifestations

72.4.1 Neurological, Psychiatric, and Behavioral Effects 72.4.2 Other Effects 72.4.3 Biomonitoring Occupational Exposures

1516

1527

Introduction Mode of Action Basic Physical and Chemical Properties Impurities Metabolism of Malathion Human Exposure to Malathion Effects seen in Animal Toxicology Studies

1541 1542

72.1

73.

Malathion: A Review of Toxicology

1540

Conclusions Acknowledgments

1495 1497

72. 70.

1536 1537 1538

73.3

Application of the Constructed PBPK/PD Model 73.3.1 Toxicity Study of Carbofuran 73.3.2 Construction and Application of a Cumulative PBPK/PD Model of Three NMCS

(c) 2011 Elsevier Inc. All Rights Reserved.

1598 1598

1599

Contents of Volume 2

x

73.4 73.5

74.

Advantages and Weaknesses of the Cumulative PBPK/PD Strategy Future Needs for the Application of PBPK Modeling in Risk Assessment Conclusion Acknowledgments

Toxicological Profile of Carbaryl 74.1 74.2 74.3

Introduction Description, Use, and Biological Mode of Action Hazard Characterization 74.3.1 Acute Toxicity 74.3.2 Subacute/Subchronic Toxicity 74.3.3 Neurotoxicity 74.3.4 Developmental and Reproductive Toxicity 74.3.5 Genotoxicity 74.3.6 Carbaryl Chronic Toxicity and Carcinogenicity Studies 74.3.7 Rat Metabolism and Toxicokinetics

Conclusion

75.

75.1 75.2 75.3

75.5

1607 1607 1607 1607 1607 1608 1608 1609 1610 1610 1612

1619

Introduction 1619 Description, Use, and Biological Mode of Action 1619 Hazard Characterization: Derivation of the Reference Dose (Rfd) 1619 75.3.1 Summary of Aldicarb Toxicity 75.3.2 Acute Toxicity 75.3.3 Neurotoxicity 75.3.4 Developmental and Reproductive Toxicity 75.3.5 Sensitivity to Infants and Children 75.3.6 Genotoxicity 75.3.7 Chronic Toxicity and Carcinogenicity 75.3.8 Human Volunteer Studies 75.3.9 Pharmacokinetics of Cholinesterase Inhibition 75.3.10 Critical Effect and Point of Departure

75.4

1603 1603 1603

1616

Aldicarb: Toxicity, Exposure and Risks to Humans

75.4.2 Summary of Residues in Food 75.4.3 Water Consumption Data 75.4.4 Summary of Residues in Drinking Water 75.4.5 Modeling of Exposure Scenarios

1602

1619 1620 1620 1621 1621 1621 1622 1622 1623 1624

Exposure Characterization

1624

75.4.1 Food Consumption Data

1624

Risk Characterization

1625 1625 1625 1626

1627

75.5.1 Acute Dietary (Food-Only) Risk Assessment 75.5.2 Acute Dietary (Water Only) Risk Assessment 75.5.3 Aggregate (Food Plus Water) Risk Assessment 75.5.4 Further Risk Characterization

1628 1629

Conclusion

1630

1627 1628

Section XI Pyrethrins and Pyrethroids 76.

Pyrethroid Chemistry and Metabolism 76.1 76.2 76.3 76.4 76.5 76.6 76.7 76.8 76.9 76.10 76.11 76.12 76.13 76.14 76.15 76.16 76.17 76.18 76.19 76.20 76.21 76.22 76.23 76.24 76.25 76.26

Introduction Allethrin (Bioallethrin, d-Allethrin, S-Bioallethrin) Bifenthrin Cycloprothrin Cyfluthrin (β-Cyfluthrin) Cyhalothrin (λ-Cyhalothrin) Cypermethrin (α-, β-, θ-, ζCypermethrin) Cyphenothrin Deltamethrin Empenthrin Etofenprox Fenpropathrin Fenvalerate (Esfenvalerate) Flucythrinate Flumethrin T-Fluvalinate (Fluvalinate) Imiprothrin Kadethrin (RU15525) Metofluthrin Permethrin Phenothrin (d-Phenothrin) Prallethrin Pyrethrins Resmethrin (Bioresmethrin, Cismethrin) Tetramethrin (d-teramethrin) Tralomethtin Conclusion

(c) 2011 Elsevier Inc. All Rights Reserved.

1635 1635 1636 1636 1638 1638 1639 1640 1641 1642 1644 1644 1645 1646 1648 1649 1649 1650 1651 1652 1653 1654 1655 1656 1657 1657 1659 1659

Contents of Volume 2

77.

xi

Toxicology and Mode of Action of Pyrethroid Insecticides 1665 77.1 77.2

Introduction Chemistry and Insecticidal Action 77.2.1 Development of Synthetic Pyrethroids 77.2.2 Structure–Activity Relationships 77.2.3 Mechanism of Insecticidal Activity

77.3

Acute Neurotoxic Actions in Mammals 77.3.1 Acute Toxicity 77.3.2 Structure–Toxicity Relationships 77.3.3 Two Syndromes of Pyrethroid Intoxication 77.3.4 Neurochemical Consequences of Pyrethroid Intoxication 77.3.5 Age-Related Differences in Pyrethroid Sensitivity 77.3.6 Reports of Neurotoxic Effects in Humans

77.4

77.5 77.6 77.7

1668 1668 1669 1669 1670

78.4 1670 1670

1671

77.8.1 Actions of Pyrethroids on Voltage-Gated Calcium Channels 77.8.2 Actions of Pyrethroids on Voltage-Gated Chloride Channels 77.8.3 Actions of Pyrethroids on GABAA Receptors

Conclusion

Chemistry and Formulations Uses Hazard Identification 78.3.1 Pharmacokinetics and Metabolism 78.3.2 Acute Toxicity 78.3.3 Short- and Long-Term Toxicity and Oncogenic Potential 78.3.4 Reproductive and Developmental Toxicity 78.3.5 Neurotoxicity 78.3.6 Metabolites other than ETU 78.3.7 Hazard Characterization

1667

1671

Actions of Pyrethroids on Other Neuronal Targets

Dialkyldithiocarbamates (EBDCs) 1689 78.1 78.2 78.3

1666

Behavioral Neurotoxicity

Neurotoxic Effects Following Dermal Exposure Developmental Neurotoxicity Actions on Voltage-Gated Sodium Channels

78.

1666

77.4.1 Effects on Motor Activity 77.4.2 Effects on Acoustic Startle Response 77.4.3 Effects on Conditioned Behavior 77.4.4 Results of Regulatory Neurotoxicity Studies

77.7.1 Electrophysiological Studies 77.7.2 Differential Sensitivity of Sodium Channel Isoforms 77.7.3 State-Dependent Actions of Pyrethroids 77.7.4 The Pyrethroid Receptor on Sodium Channels 77.7.5 Correlation of Sodium Channel Effects with Toxicity

77.8

1665 1666

Section XII Herbicides

1671 1672

78.5 78.6

1672

1672 1673

Dose–Response

1689 1689 1689 1690 1691 1691 1699 1700 1700 1700

1700

78.4.1 NOAEL and Acceptable Daily Intake—ETU 78.4.2 NOAEL and Acceptable Daily Intake—EBDCS 78.4.3 Acute Reference Dose 78.4.4 Endpoints for Assessment of Dermal and Respiratory Exposure 78.4.5 Carcinogenicity Classification and Low Dose Risk Assessment

1702

Toxicology in Humans Risk Characterization

1703 1703

78.6.1 Dietary Exposure and Risks 78.6.2 Worker Exposure and Risks

1703 1704

Conclusion

1704

1700 1701 1702

1702

1673 1673 1674 1675

79.

Symmetrical Triazine Herbicides: A Review of Regulatory Toxicity Endpoints 1711 79.1

1675 1677

79.2

1677

1677

1679

79.3

1680

1682

79.4

Introduction

1711

79.1.1 Chemistry 79.1.2 Uses

1711 1711

Hazard Identification

1711

79.2.1 Acute Studies 79.2.2 Toxicity after Repeat Exposure 79.2.3 Developmental and Reproductive Toxicity 79.2.4 Mutagenicity 79.2.5 Oncogenicity Assessment

1713 1713 1714 1716 1716

Mode of Action for Mammary Tumor Formation in the Sprague-Dawley Rat at High Doses

1717

Epidemiology

1719

(c) 2011 Elsevier Inc. All Rights Reserved.

Contents of Volume 2

xii

79.5 79.6

80.

Chlorotriazine Cancer Classification 1720 Overall Hazard Assessment 1721

Phenylurea Herbicides 80.1 80.2 80.3 80.4

82.4

82.5

1725

Introduction Diuron Fluometuron Isoproturon Conclusion

1725 1725 1727 1728 1730

82.6

82.7

81.

Protoporphyrinogen Oxidase-Inhibiting Herbicides 81.1 81.2

1733

Introduction Commercially Available Protox Inhibitors

1733

81.2.1 Diphenyl Ether Protoporphyrinogen Oxidase Inhibitors 81.2.2 Non–Diphenyl Ether Protoporphyrinogen Oxidase Inhibitors

81.3

81.4

81.5

81.6

82.

82.3

83. 1733

1734

81.3.1 Crops and Weeds 81.3.2 Mode of Application

1734 1736

Behavior in Plants

1737

81.4.1 Absorption, Translocation, and Metabolism 81.4.2 Mode of Action 81.4.3 Mode of Resistance 81.4.4 Genetically Engineered Resistance

1742

Environmental Impact

1742

81.5.1 Interaction with Soil 81.5.2 Degradation in the Environment 81.5.3 Ecotoxicology

1742

1737 1738 1740

1742 1742

Mammalian Toxicology

1743

81.6.1 Skin and Oral 81.6.2 Teratogenicity and Mutagenicity 81.6.3 Effects on Mammalian Porphyrin Metabolism 81.6.4 Metabolic Degradation in Animals

1743 1745

1746

Conclusion

1747

1745

1760 1760 1760

Metolachlor

1762

82.5.1 Identity, Properties, and Uses 82.5.2 Toxicity to Laboratory Animals

1762 1762

Propachlor

1763

82.6.1 Identity, Properties, and Uses 82.6.2 Toxicity to Laboratory Animals 82.6.3 Human Experience

1763 1763 1765

Mode-of-Action Evaluations: Oncogenicity

1765

82.7.1 Rat Nasal Tumors 82.7.2 Rat Stomach Tumors 82.7.3 Rat Thyroid Tumors

1765 1767 1767

Common Mechanism of Toxicity

1768

Paraquat 83.1

1771

Identity, Properties, and Use

1771

83.1.1 83.1.2 83.1.3 83.1.4

1771 1771 1771

Chemical Name Structure Synonyms Physical and Chemical Properties 83.1.5 History, Formulations, and Uses

1733

Agricultural Use

83.2

1771 1771

Toxicity to Laboratory Animals

1772

83.2.1 83.2.2 83.2.3 83.2.4 83.2.5

1772 1772 1774 1774

83.2.6

83.2.7 83.2.8 83.2.9 83.2.10 83.2.11 83.2.12 83.2.13 83.2.14

1753

83.2.15 83.2.16

Introduction Alachlor

1753 1753

83.2.17

82.2.1 Identity, Properties, and Uses 82.2.2 Toxicity to Laboratory Animals 82.2.3 Human Experience

1753 1753 1757

Chloracetanilides 82.1 82.2

82.8

1733

Butachlor 82.4.1 Identity, Properties, and Uses 82.4.2 Toxicity to Laboratory Animals

Acetochlor

1757

83.2.18 83.2.19 83.2.20

82.3.1 Identity, Properties, and Uses 82.3.2 Toxicity to Laboratory Animals

1757 1758

83.2.21

(c) 2011 Elsevier Inc. All Rights Reserved.

Signs of Toxicity Acute Toxicity Irritation and Sensitization Subchronic Toxicity Mutagenic and Carcinogenic Potential Effects on Reproduction, Embryotoxicity and Teratogenicity Pathology of the Lung Absorption Distribution Metabolism Excretion Accumulation of Paraquat into the Lung Efflux of Paraquat from the Lung Biochemical Mechanisms of Paraquat Toxicity Lipid Peroxidation Hypothesis Oxidation of NADPH Hypothesis The Role of Mitochondria in the Toxicity The Involvement of Oxygen Effects on the Kidney Effects on the Central Nervous System Effects on Other Organs

1775

1776 1776 1777 1778 1779 1780 1781 1782 1783 1785 1786 1786 1787 1787 1788 1791

Contents of Volume 2

83.3

84.

83.2.22 Treatment of Poisoning in Animals 83.2.23 Adsorption from the Gastrointestinal Tract 83.2.24 Removal from the Bloodstream 83.2.25 Prevention of Accumulation into the Lung 83.2.26 Free Radical Scavenging 83.2.27 Prevention of Lung Fibrosis

1792 1792 1793

Toxicity to Humans

1793

83.3.1 Experimental Exposure 83.3.2 Accidental and Intentional Poisoning 83.3.3 Use Experience 83.3.4 Atypical Cases of Various Origins 83.3.5 Clinical Findings and Dosage Response 83.3.6 Laboratory Findings 83.3.7 Absorption 83.3.8 Distribution 83.3.9 Metabolism 83.3.10 Excretion 83.3.11 Pathology 83.3.12 Treatment of Poisoning

1793

Phenoxy Herbicides (2,4-D) 84.1 84.2

84.3 84.4 84.5 84.6

84.7 84.8

85.

xiii

86.3

1829

1829 1830 1830 1831 1831 1831 1831

History of Use Formulations Human Exposure to 2,4-D Toxicological Studies

1831 1832 1832 1832

84.6.1 84.6.2 84.6.3 84.6.4 84.6.5 84.6.6 84.6.7

1832 1833 1833 1833 1833 1834 1840

Studies in Humans Summary

86.2

1800 1804 1805 1805 1806 1806 1806 1807

87.

Synonyms Physical and Chemical Properties Formulations and Uses

1849 1849 1849

1849 1850 1850 1850 1850 1851 1851

1853

Identity, Properties, and Uses

1853

86.1.1 Chemical Names 86.1.2 Physical and Chemical Properties 86.1.3 Structure 86.1.4 History and Uses

1853 1853 1853 1854

Toxicity to Laboratory Animals

1854

86.2.1 Basic Findings 86.2.2 Absorption, Distribution, Metabolism, and Excretion 86.2.3 Effects on Organs and Tissues 86.2.4 Effects on Reproduction 86.2.5 Pathology 86.2.6 Genotoxicity Studies

1854 1861 1862 1862 1862 1863

Toxicity to Humans

1863

86.3.1 Use Experience 86.3.2 Treatment of Poisoning

1863 1863

Toxicology of Triazolopyrimidine Herbicides 1865 87.1 87.2

87.3

Introduction Cloransulam-Methyl

1865 1866

87.2.1 Identity, Properties, and Uses 87.2.2 Toxicity to Laboratory Animals 87.2.3 Toxicity to Humans

1866 1866 1868

Diclosulam

1868

87.3.1 Identity, Properties, and Uses 87.3.2 Toxicity to Laboratory Animals 87.3.3 Toxicity to Humans

87.4

87.5

1840 1841

1849

Toxicokinetics Toxicity to Laboratory Animals Toxicity to Humans Reproductive Effects Genotoxic Effects Treatment of Poisoning Conclusion

Imidazolinones 86.1

1798

84.2.1 84.2.2 84.2.3 84.2.4 84.2.5 84.2.6 84.2.7

Absorption Distribution Pharmacokinetics Metabolism Excretion Animal Studies Genotoxicity

86.

1794 1795

1829 1829

Dicamba 85.1 85.2 85.3

1791 1792

Introduction Physical and Chemical Properties 2,4-D Acid, Salts, and Esters 2,4,5-T 2,4-DB Dichlorprop (2,4-DP) Mecoprop (MCPP) Mcpa Silvex

85.4 85.5 85.6 85.7 85.8 85.9

1791

87.6

87.7

1868 1869 1870

Florasulam

1870

87.4.1 Identity, Properties, and Uses 87.4.2 Toxicity to Laboratory Animals 87.4.3 Toxicity to Humans

1870 1871 1873

Flumetsulam

1873

87.5.1 Identity, Properties, and Uses 87.5.2 Toxicity to Laboratory Animals 87.5.3 Toxicity to Humans

1873 1873 1875

Metosulam

1875

87.6.1 Identity, Properties, and Uses 87.6.2 Toxicity to Laboratory Animals 87.6.3 Toxicity to Humans

1875 1875 1877

Penoxsulam

1877

(c) 2011 Elsevier Inc. All Rights Reserved.

Contents of Volume 2

xiv

87.8

87.7.1 Identity, Properties, and Uses 87.7.2 Toxicity to Laboratory Animals 87.7.3 Toxicity to Humans

1877 1878 1879

Pyroxsulam

1879

87.8.1 Identity, Properties, and Uses 87.8.2 Toxicity to Laboratory Animals 87.8.3 Toxicity to Humans

1879 1880 1881

Conclusion

1881

89.4.4 Toxicokinetics and Metabolism in Laying Hens 1905

89.5

Section XIII Fungicides 88.

A Toxicological Assessment of Sulfur as a Pesticide 88.1

88.2

Introduction

1889

88.1.1 Usage 88.1.2 Environmental Fate

1889 1889

88.6

89.

89.3 89.4

1890 1890 1891 1891 1892

1892 1893 1894

88.5.1 Occupational exposure

1895

Discussion Conclusion Acknowledgments

1899 1900 1900

90.1

90.2

89.2.1 Chemical Name 89.2.2 Synonyms 89.2.3 Physical and Chemical Properties 89.2.4 History, Formulations, and Uses

1903 1903 1904 1904

90.4

1904 1904

89.4.1 Toxicokinetics in Rats 1904 89.4.2 Metabolic Pathways in Rats 1904 89.4.3 Toxicokinetics and Metabolism in Lactating Goats 1905

1909 1910 1911

Mutagenicity Toxicity to Humans

1911 1911

89.7.1 Direct Observations and Health Records 89.7.2 Diagnosis of Poisoning 89.7.3 Sensitization Observations 89.7.4 Proposed Treatment

1911 1911 1912 1912

Conclusion

1912

90.5

1907

1907 1908

1915

Introduction

1915

90.1.1 Overview 90.1.2 History and Use 90.1.3 Toxicological Overview

1915 1916 1916

Physical Properties and Chemical Reactions

1917

90.2.1 90.2.2 90.2.3 90.2.4 90.2.5

1917 1918 1918 1920 1921

Overview Physical Properties Chemical Reactions Metabolism Summary

Toxicology

1922

90.3.1 90.3.2 90.3.3 90.3.4

1922 1924 1926

Acute Toxicology Subchronic Toxicity Chronic Toxicity Developmental and Reproductive Toxicity 90.3.5 Mutagenicity 90.3.6 Carcinogenicity

1903 1903 1903

1906 1906 1906

Captan and Folpet

90.3

Introduction Identity, Properties, and Uses

Biological Mode of Action Absorption, Distribution, metabolism, and Excretion

90.

1892

Toxicology of Sulfur Dioxide Veterinary Effects of Sulfur Human Health Effects of Sulfur

Cyprodinil: A Fungicide of the Anilinopyrimidine Class 89.1 89.2

89.6 89.7

Toxicology Profile of Elemental Sulfur 1890 88.2.1 Acute Exposure Oral Toxicity: 81-1 88.2.2 Acute Exposure Dermal Toxicity: 81-2 88.2.3 Acute Exposure Inhalation Toxicity: 81-3 88.2.4 Primary Eye Irritation: 81-4 88.2.5 Primary Dermal Irritation: 81-5 88.2.6 Primary Dermal Sensitization: 81-6

88.3 88.4 88.5

1889

Toxicity to Laboratory Animals 89.5.1 Acute Toxicity 89.5.2 Subchronic Toxicity 89.5.3 Chronic Toxicity and Oncogenicity 89.5.4 Effects on Liver Xenobiotic Metabolizing Enzymes in the Rat 89.5.5 Effects on Liver and Plasma Lipids in the Rat 89.5.6 Effects on Reproduction and Development 89.5.7 Neurotoxic Effects 89.5.8 Pharmacological Effects

1927 1928 1933

Common Mechanism of Toxicity

1937

90.4.1 Captan and Folpet 90.4.2 Captafol 90.4.3 Dichlofluanid and Tolylfluanid

1937 1938 1938

Human Risk Assessment

1939

90.5.1 Cancer 90.5.2 Noncancer

1939 1939

Conclusion Acknowledgments

1940 1941

(c) 2011 Elsevier Inc. All Rights Reserved.

Contents of Volume 2

91.

xv

Mammalian Toxicokinetics and Toxicity of Chlorothalonil 91.1

91.2

91.3

91.4

91.5

Identity and Uses of Chlorothalonil

1951

91.1.1 Physical and Chemical Properties

1951

Mammalian Toxicokinetics

1951

91.2.1 Oral Administration 91.2.2 Dermal Administration

1951 1953

Acute toxicity

1953

91.3.1 91.3.2 91.3.3 91.3.4 91.3.5 91.3.6 91.3.7

1953 1954 1954 1954 1954 1954 1955

Oral Dermal Intraperitoneal Inhalation Skin Irritation Eye Irritation Summary of Acute Toxicity

Sensitization

1955

91.4.1 Skin Sensitization 91.4.2 Respiratory Sensitization

1955 1955

1955

91.5.1 Oral 91.5.2 Dermal 91.5.3 Inhalation

1955 1956 1957

Chronic Toxicity Genotoxicity

1957 1958

91.7.1 In Vitro Genotoxicity Studies 91.7.2 In Vivo Genotoxicity Studies

1958 1958

91.8

Carcinogenicity

1958

91.8.1 Mode of Carcinogenic Action

1959

91.9

Reproductive Toxicity

1960

91.9.1 Developmental Toxicity 91.9.2 Fertility

1960 1961

91.10 Investigative Toxicity Studies 91.10.1 Acute Effects on Hepatic and Renal Glutathione Content 91.10.2 Effect of Dietary vs Gavage Dosing on Renal Toxicity in the Rat

93.3

93.4

93.5

93.6

93.7

1962

94.

94.1

94.2

1967

Introduction Glyphosate

1967 1967

92.2.1 Identity, Properties, and Uses 92.2.2 Toxicity to Laboratory Animals 92.2.3 Human Experience

1967 1968 1970

1975 1975

93.2.1 Identity, Properties, and Uses 93.2.2 Formulations and Production 93.2.3 Toxicity to Laboratory Animals 93.2.4 Toxicity to Humans

1976 1976

94.3

1977 1993

TDE

2004

93.3.1 Identity, Properties, and Uses 93.3.2 Toxicity to Laboratory Animals 93.3.3 Toxicity to Humans

2004 2004 2007

Ethylan

2008

93.4.1 Identity, Properties, and Uses 93.4.2 Toxicity to Laboratory Animals 93.4.3 Toxicity to Humans

2008 2008 2009

Methoxychlor

2010

93.5.1 Identity, Properties, and Uses 93.5.2 Toxicity to Laboratory Animals 93.5.3 Toxicity to Humans

2010 2010 2014

Chlorobenzilate

2014

93.6.1 Identity, Properties, and Uses 93.6.2 Toxicity to Laboratory Animals 93.6.3 Toxicity to Humans

2014 2015 2015

Dicofol

2015

93.7.1 Identity, Properties, and Uses 93.7.2 Toxicity to Laboratory Animals 93.7.3 Toxicity to Humans

2015 2015 2016

2016

93.8.1 Identity, Properties and Uses 93.8.2 Toxicity

2016 2016

Conclusion

2016

Boric Acid and Inorganic Borate Pesticides

1963

1975

Introduction DDT

93.8. Acetofenate

1961 1962 1962 1962

92.1 92.2

93.1 93.2

1961

91.11.1 91.11.2 91.11.3 91.11.4

Inhibitors of Aromatic Acid Biosynthesis

Toxicology of DDT and Some Analogues

1961

91.11 Human Data Dermal Effects Ocular Effects Respiratory Effects Clinical Cases and Poisoning Incidents

Section XIV Other Selected Pesticides 93.

Subchronic toxicity

91.6 91.7

92.

1951

2033

Introduction

2033

94.1.1 Background 94.1.2 Chapter Coverage 94.1.3 Product Uses

2033 2033 2034

Names and Chemical/Physical properities

2035

94.2.1 Boric Acid 94.2.2 Sodium Borate Salts

2035 2035

Exposure

2035

94.3.1 Dietary Exposure 94.3.2 Occupational Exposure 94.3.3 Residential Exposure

2035 2035 2035

(c) 2011 Elsevier Inc. All Rights Reserved.

Contents of Volume 2

xvi

94.4

94.5

94.6

94.3.4 Environmental Exposure

2036

Biological Importance

2036

94.4.1 Essentiality in Plants 94.4.2 Biological Importance in Animals (Vertebrates) 94.4.3 Biological Importance in Humans

2036 2037

Interactions with the GammaAminobutyric Acid A-Receptor: Polychlorocycloalkanes and Recent Congeners and Other Ligands 2065

2037

96.1 96.2

Toxicokinetics

2038

94.5.1 94.5.2 94.5.3 94.5.4

2038 2038 2039 2039

Absorption Distribution Metabolism Excretion

Toxicology

96.

2040

94.8

95.

Accidental Poisionings

2042

94.7.1 Animals 94.7.2 Humans

2042 2042

Findings from Studies with Humans

2043

94.8.1 Occupational Studies 94.8.2 Epidemiological Studies

2043 2044

Imidacloprid: A Neonicotinoid Insecticide 95.1 95.2

95.3 95.4 95.5 95.6

95.7

2055 2055

95.2.1 Chemistry 95.2.2 Nicotinic Activity

2055 2056

Metabolism and Toxicokinetics Mammalian Toxicology Acute Toxicity Subchronic Toxicity

2057 2057 2057 2058

95.6.1 Rat 95.6.2 Dog

2058 2059

Chronic Toxicity and Carcinogenicity

2059

95.7.1 Rat 95.7.2 Mouse 95.7.3 Classification for Carcinogenicity 95.7.4 Dog

2059 2059

Mutagenicity Developmental Toxicity

2060 2060

95.9.1 Rat 95.9.2 Rabbit

2060 2060

95.10 Reproductive Toxicity 95.11 Neurotoxicity

2061 2061

95.8 95.9

95.11.1 95.11.2 95.11.3 95.11.4

96.3

96.4

2061 2061 2061

Conclusion

2063

2062 2062

2066

2071 2073 2075

Structure–Toxicity Relationship and Mode of Action

2076

Molecular Mechanism of Action

2082

96.4.1 Topography of The GammaAminobutyric Acid A-Receptor 96.4.2 Molecular Biology of Cyclodiene Resistance 96.4.3 Molecular Toxicology of Noncompetitive Chloride Ionophore Blockers

2082 2083

2083

Conclusion

2060 2060

General Acute Neurotoxicity Subchronic Neurotoxicity Comparison with other Neonicotinoids 95.11.5 Developmental Neurotoxicity

2065 2065

96.3.1 Fully Chlorinated Cyclodienes: Substituted Hexachloronorbornenes (HCNB) 2076 96.3.2 Compounds with Fewer, or no Chlorine Atoms 2078 96.3.3 Links between Polychlorocycloalkane and Recent Heterocyclics Apparently Acting at the Chloride Ionophore 2081

2055

Introduction Historical Overview

2065

96.2.1 Background 96.2.2 Lindane, Aldrin, Dieldrin, Isodrin, and Endrin and Analogues 96.2.3 Heptachlor, Chlordene, Dihydroheptachlor, Chlordane, and Isobenzan 96.2.4 Endosulfan (Thiodan) 96.2.5 Toxaphene, Mirex, Chlordecone (Kepone)

94.6.1 Efficacy (Invertebrate and Fungi) 2040 94.6.2 Laboratory Studies 2040

94.7

Introduction Discovery of Polychlorocycloalkane Metabolism as a Factor in Toxicity

97.

The Role of P-glycoprotein in Preventing Developmental and Neurotoxicity: Avermectins – A Case Study 97.1 97.2 97.3 97.4 97.5 97.6 97.7

2088

2093

Introduction Chemistry and Formulations Uses Mode of Action of the Avermectins Hazard Identification and Dose Response Humans: Experience with Ivermectin Risk Characterization

(c) 2011 Elsevier Inc. All Rights Reserved.

2093 2093 2095 2096 2096 2101 2102

Contents of Volume 2

97.8

98.

xvii

DEET 98.1 98.2 98.3

99.5.2 Non-Dietary Exposure

Importance of the p-Glycoprotein Blood–Brain Barrier Conclusion

2104 2107

2111 Introduction Chemistry Overview of Toxicology Studies

2111 2111 2111

98.3.1 98.3.2 98.3.3 98.3.4 98.3.5

2112 2113 2114 2115

Acute Toxicity Studies Subchronic Toxicity Studies Developmental Toxicity Reproductive Toxicity Chronic Toxicity and Oncogenicity 98.3.6 Neurotoxicity 98.3.7 Genotoxicity Studies

98.4

The Safety Assessment of Piperonyl Butoxide 99.1 99.2 99.3

99.4 99.5

2146

99.6.1 Cancer 99.6.2 NonCancer Effects

2146 2146

100. Rodenticides 100.1 Introduction 100.2 Fluoroacetic Acid and its Derivatives 100.2.1 Sodium Fluoroacetate 100.2.2 Fluoroacetamide 100.2.3 Fluoroethanol 100.3.1 Pyriminil

100.4 Thioureas 100.4.1 Antu

100.5 Anti-Vitamin K Compounds

Pharmacokinetic Studies: Animals and Humans

2117

2117 2118

2119

2120 2121 2121 2122

2127

Chemistry and Formulations Uses Hazard Identification

2127 2127 2128

99.3.1 99.3.2 99.3.3 99.3.4 99.3.5 99.3.6 99.3.7

Acute Toxicity Subchronic Toxicity Reproductive Toxicity Developmental Toxicity Chronic Toxicity/Oncogenicity Genotoxicity Mode of Action Considerations for Oncogenicity 99.3.8 Human Studies 99.3.9 Human Experience

2128 2128 2131 2134 2136 2140

Pharmacodynamics Exposure Assessment

2144 2145

99.5.1 Dietary Exposure

2145

2140 2143 2144

2145

Risk characterization

100.3 Substituted Ureas 2115 2116 2117

98.4.1 Absorption, Distribution, Metabolism, and Excretion in Rats 98.4.2 Absorption, Metabolism, and Excretion in Humans 98.4.3 DEET and Interactions with Other Environmental Chemicals (Ethanol and Sunscreens) 98.4.4 DEET and Interactions with Other Environmental Chemicals 98.4.5 Human Aspects: Clinical Case Reports (Dermal Reactions) 98.4.6 Clinical Case Reports: Adverse Neurological Effects 98.4.7 Regulatory Risk Assessment

99.

99.6

100.5.1 Overview 100.5.2 Resistance to Anticoagulant Rodenticides 100.5.3 Warfarin 100.5.4 Coumafuryl 100.5.5 Diphacinone 100.5.6 Brodifacoum 100.5.7 Chlorophacinone 100.5.8 Difenacoum 100.5.9 Bromadiolone 100.5.10 Difethialone

100.6 Vitamin D-Related Compounds 100.6.1 Ergocalciferol 100.6.2 Cholecalciferol

100.7 Miscellaneous Synthetic Organic Rodenticides 100.7.1 100.7.2 100.7.3 100.7.4 100.7.5

Chloralose Norbormide Bromethalin Banned Compounds On-going Research for New Chemical Families with Rodenticidal Properties

Conclusion

2153 2153 2154 2154 2160 2163

2163 2164

2167 2168

2171 2171 2173 2174 2182 2182 2184 2188 2190 2192 2193

2194 2194 2197

2200 2200 2202 2204 2205

2205

2206

101. Toxicology and Safety Evaluation of the New Insect Repellent Picaridin (Saltidin) 2219 101.1 Introduction 101.2 General Overview 101.2.1 Chemistry 101.2.2 Mode of Action 101.2.3 Effectiveness Against Disease Vectors

101.3 Metabolism and Toxicokinetics

(c) 2011 Elsevier Inc. All Rights Reserved.

2219 2219 2219 2219 2220

2221

Contents of Volume 2

xviii

101.4 Mammalian Toxicology

2223

101.4.1 Acute Toxicity 101.4.2 Subchronic toxicity 101.4.3 Chronic Toxicity and Oncogenicity 101.4.4 Genotoxicity

101.5 Developmental Toxicity

2223 2224 2224 2225 2225 2225

101.6 Reproductive Toxicity 101.7 Neurotoxicity

2225 2226

101.9

102.8 Reference Values and Conclusions Acknowledgments

2240 2241 2241

2241 2242

2225

101.5.1 Rat 101.5.2 Rabbit

101.8

102.7.2 Diagnosis of Poisoning 102.7.3 Sensitization Observations 102.7.4 Proposed Treatment

101.7.1 Acute 101.7.2 Subchronic

2226 2226

Dermal Absorption

2226

101.8.1 Rat 101.8.2 Human Volunteers 101.8.3 Human Skin

2226 2226 2226

Exposure and Safety Evaluation Conclusion

2227 2227

Section XV Fumigants 103. Sulfuryl Fluoride 103.1 Chemistry and Formulations 103.2 Uses 103.3 Hazard Identification: Toxicity to Laboratory Animals (Pre-1980)

2245 2245

103.3.1 Acute Toxicity 103.3.2 Subchronic Toxicity

2246 2246

103.4 Toxicity to Laboratory Animals (Post-1980)

102. Chlorantraniliprole: An Insecticide of the Anthranilic Diamide Class

2231

102.1 Introduction 102.2 Identity, Properties, and Uses

2231 2231

102.2.1 Chemical Name 102.2.2 Synonyms 102.2.3 Physical and Chemical Properties 102.2.4 History, Formulations, and Uses

2231 2231

102.3 Biological Mode of Action 102.4 Absorption, Distribution, Metabolism, and Excretion 102.4.1 Toxicokinetics in Rats 102.4.2 Metabolic Pathways in Rats 102.4.3 Plasma Concentrations of Parent and Metabolites in Feeding Studies

102.5 Toxicity to Laboratory Animals 102.5.1 Acute Toxicity 102.5.2 Subchronic Toxicity 102.5.3 Chronic Toxicity and Oncogenicity 102.5.4 Effects on Adrenal Function 102.5.5 Effects on Reproduction and Development 102.5.6 Neurotoxic Effects 102.5.7 Immunotoxic Effects

102.6 Genotoxicity 102.7 Toxicity to Humans

2231 2232

2232 2232 2232 2233

Acute Toxicity Acute Neurotoxicity Time to Incapacitation Therapeutic/Amelioration of Toxicity 103.4.5 Repeated Exposures 103.4.6 Subchronic Toxicity 103.4.7 Chronic Toxicity 103.4.8 Teratology Studies 103.4.9 Reproduction Toxicity 103.4.10 Genetic Toxicity 103.4.11 Uptake and Metabolism

103.5 Toxicology in Humans 103.6 Conclusion 103.6.1 Risk Characterization

104. Phosphine 104.1.1 Physical Properties 104.1.2 Chemistry

2246

2246 2246 2247 2247 2247 2248 2248 2252 2254 2254 2255 2255

2256 2256 2256

2259 2259 2259 2259

2233

104.2 Sources, Uses, and Formulations

2260

2233 2235

104.2.1 Natural Sources 104.2.2 Commercial Sources

2260 2260

104.3 Toxicology 2237 2237 2238 2239 2240

2240 2240

102.7.1 Direct Observations and Health Records

103.4.1 103.4.2 103.4.3 103.4.4

104.1 Identity, Properties, and Uses 2233

2245

2240

104.3.1 Overview

2260 2260

104.4 Toxicity and Mode of Action

2260

104.4.1 Acute Toxicity

2260

104.5 Animal Dose/Response 104.5.1 Threshold for Lethality 104.5.2 Acute and Subacute Dose/Response

104.6 Absorption, Distribution, Metabolism, and Excretion 104.7 Cellular and Molecular Studies

(c) 2011 Elsevier Inc. All Rights Reserved.

2261 2261 2261

2261 2261

Contents of Volume 2

xix

104.7.1 General 104.7.2 Cytochromes and Cytochrome Oxidase 104.7.3 Hemoglobin

104.8 Peroxidases, Lipid Peroxidation, Catalase, and Cholinesterase 104.9 Genotoxicity, Cancer and Reproductive Effects 104.10 Treatment of Poisoning 104.11 Regulatory Notes (Exposure Guidelines) 104.12 Summary and Comments

105. Methyl Bromide

105.3.5 105.3.6 105.3.7 105.3.8

Acute Toxicity Subchronic Toxicity Genetic Toxicity Developmental and Reproductive Toxicity Chronic Toxicity and Oncogenicity: Inhalation Chronic Toxicity and Oncogenicity: Dietary Neurotoxicity Specific Target Organ Effects

105.4 Metabolism

2262 2262

106.3.1 106.3.2 106.3.3 106.3.4

Acute Toxicity Repeated Dose Toxicity Effects on Reproduction Absorption, Distribution, Metabolism, and Excretion 106.3.5 Short-Term Assays 106.3.6 Chronic Toxicity and Oncogenicity Assays

2262 2263 2263 2264 2264

106.4 Dose Response 106.5 Toxicology in Humans

105.5 Human Exposure Conclusion

106.5.1 Experimental Exposure 106.5.2 Accidental Poisoning 106.5.3 Use Experience

2267

106.6 Summary Risk Characterization

2267 2268 2268 2269 2270

107. Metam-Sodium 107.1 107.2 107.3 107.4 107.5

2272 2273

107.6 107.7 107.8 107.9

2273 2274

2275 2276

108. Methyl Iodide 108.1 108.2 108.3 108.4

2276 2277

2277 2277

106. 1,3-Dichloropropene Chemical Name Structure Synonyms Physical and Chemical Properties

2281 2282 2282 2282 2283 2283 2285 2285

2286 2288 2288 2288 2288

2289

2293 2293 2294 2294 2295 2297 2298 2302 2303 2304

2281 2281 2281 2281 2281

2307

Background Methyl Iodide Properties Registration Issues Exposure Issues

2307 2307 2308 2309

108.4.1 Quantifying Exposure of Plant Pests to MeI

2309

108.5 Crop Production 108.6 Environmental Fate of Methyl Iodide as a Soil Fumigant

2281

106.1 Chemistry and Formulations

Introduction Acute toxicity Subchronic Toxicity Genetic Toxicity Developmental and Reproductive Toxicity Chronic/Oncogenicity Toxicity Nurotoxicity Other Studies (Mammalian) Metabolism

2281

2275

2275

105.4.1 Absorption 105.4.2 Distribution 105.4.3 Identification and Quantitation of Metabolites 105.4.4 Excretion

106.1.1 106.1.2 106.1.3 106.1.4

106.2 Uses 106.3 Hazard Identification

2267

105.1 Introduction 105.2 Chemical Properties and Pesticidal Uses of Methyl Bromide 105.3 Toxicology of Methyl Bromide 105.3.1 105.3.2 105.3.3 105.3.4

106.1.5 Formulations

2261

2311 2311

108.6.1 Transformation in Aqueous Solution 108.6.2 Transformation in Soil

2311 2312

Conclusion

2316

Index

(c) 2011 Elsevier Inc. All Rights Reserved.

2319

Contributors

Husein Ajwa, Department of Plant Sciences, University of California, Davis, California 95616, USA

Ann M. Blacker, Bayer CropScience, Research Triangle Park, North Carolina 27709, USA

Iris S. Ale, Republic University of Uruguay, Montevideo, Uruguay

Jerry N. Blancazo, U.S. Environmental Protection Agency, Research Triangle Park, North Carolina 27711, USA

Sandra L. Allen, Regulatory Science Associates, Dunoon, Argyll, United Kingdom Judith Alsop, California Poison Control Sacramento, California 95817, USA

System,

Gail Arce, Griffin LLC, Valdosta, Georgia 31603, USA D. J. Ashworth, USDA-ARS United States Salinity Laboratory, Riverside, California 92507, USA Sharada Balakrishnan, University Riverside, California 92521, USA

of

California,

John B. Barnett, Department of Microbiology, Immunology and Cell Biology, West Virginia University School of Medicine, Morgantown, West Virginia 26506, USA Dana B. Barr, U.S. Department of Health and Human Services, Centers for Disease Control and Prevention, Chamblee, Georgia 30341, USA Terrell Barry, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California 95814, USA Ronald E. Baynes, College of Veterinary Medicine, North Carolina State University, Raleigh, North Carolina 27606, USA Sheryl Beauvais, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California 95814, USA Karin S. Bentley, DuPont Crop Protection, E. I. du Pont de Nemours and Company, Newark, Delaware 19713, USA Craig E. Bernard, Rio Tinto Borax, Boron, California 93516, USA Nida Besbelli, World Health Organization, European Centre for Environment and Health, 53113 Bonn, Germany Richard Billington, Dow AgroSciences Indianapolis, Indiana 46268, USA

LLC,

Charles B. Breckenridge, Syngenta Crop Protection, Inc., Greensboro, North Carolina 27419, USA Gerald T. Brooks, University of Portsmouth, Burgess Hill, United Kingdom James Bruckner, Kansas Life Sciences Innovations, Kansas City, Kansas 64108, USA Kathleen M. Brundage, Department of Microbiology, Immunology and Cell Biology, West Virginia University School of Medicine, Morgantown, West Virginia 26506, USA Quang Bui, Cerexagri, Inc., King of Prussia, Pennsylvania 19406, USA Franca M. Buratti, Istituto Superiore di Sanità, Viale Regina Elena 299, Rome, Italy James S. Bus, The Dow Chemical Company, Midland, Michigan 48674, USA Geoffrey M. Calvert, Centers for Disease Control and Prevention, Cincinnati, Ohio 45226, USA Linda L. Carlock, Toxicology and Regulatory Consulting John E. Casida, Environmental Chemistry and Toxicology Laboratory, Department of Environmental Science, Policy and Management, University of California, Berkeley, California, 94720, USA Howard W. Chambers, Department of Entomology and Plant Pathology, Mississippi State University, Mississippi State, Mississippi 39762, USA Janice E. Chambers, Center for Environmental Health Sciences, College of Veterinary Medicine, Mississippi State University, Mississippi State, Mississippi 39762, USA Heidi P. Chan, Department of Dermatology, University of California, San Francisco, California 94143, USA Graham Chester, OCCUBEX RA Limited, Hampshire, UK xxiii

(c) 2011 Elsevier Inc. All Rights Reserved.

Contributors

xxiv

J. Marshall Clark, Veterinary and Animal Sciences Department, University of Massachusetts, Amherst, Massachusetts 01003, USA

Jeffrey B. Evans, U.S. Environmental Protection Agency, Office of Pesticide Programs, Washington, D.C., 20460, USA

Thomas Class, PTRL Europe, D-89081 Ulm, Germany

Donna Farmer, Monsanto Company, St. Louis, Missouri 63198, USA

M. Scott Clifton, U.S. Environmental Protection Agency, National Exposure Research Laboratory, Research Triangle Park, North Carolina 27711, USA Roger Cochran, California Department of Pesticide Regulation, Sacramento, California 95814, USA Emma Di Consiglio, Istituto Superiore di Sanità, Viale Regina Elena 299, Rome, Italy Curtis C. Dary, U.S. Environmental Protection Agency, Las Vegas, Nevada 89119, USA Franck E. Dayan, United States Department of Agriculture, University, Mississippi 38677, USA Allison L. De Vries, Centers for Disease Control and Prevention, Cincinnati, Ohio 45226, USA Kelly J. Dix, Research Triangle Institute, Research Triangle Park, North Carolina 27709, USA Michael H. Dong, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California 95814, USA

Allan S. Felsot, Entomology and Environmental Toxicology, Washington State University, Richland, Washington 99352, USA Penelope A. Fenner-Crisp, U.S. Environmental Protection Agency (Retired) Joan L. Fletcher, DuPont Crop Protection, E. I. du Pont de Nemours and Company, Newark, Delaware 19714, USA Sara Flores, University of California, San Francisco, California 94143, USA Roy Fortmann, U.S. Environmental Protection Agency, National Exposure Research Laboratory, Research Triangle Park, North Carolina 27711, USA Toshio Fujita, EMIL Project, Kyoto, Japan Derek W. Gammon, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California 95814, USA Suduan Gao, USDA-ARS, Parlier, California 93648, USA

Timothy A. Dotson, UCB Chemicals Corporation, Smyrna, Georgia 30080, USA

V. F. Garry, University of Minnesota, Minneapolis, Minnesota 55455, USA

John Doull, Kansas Life Sciences Innovations, Kansas City, Kansas

Sean C. Gehen, Dow AgroSciences LLC, Indianapolis, Indiana 46268, USA

Jeffrey H. Driver, infoscientific.com & risksciences.net, LLC, Arlington, Virginia 22201, USA

Panos Georgopoulos, Environmental and Occupational Health Sciences Institute, a joint institute of UMDNJRW Johnson Medical School and Rutgers University, Piscataway, New Jersey, 08854, USA

Stephen O. Duke, United States Department Agriculture, University, Mississippi 38677, USA

of

M. Bigelow Dyk, University of California, Riverside, California 92521, USA David A. Eastmond, College of Natural and Agricultural Sciences, University of California, Riverside, California 92521, USA David L. Eaton, Department of Environmental and Occupational Health Sciences, University of Washington, Seattle, Washington 98105, USA Marion Ehrich, Virginia–Maryland Regional College of Veterinary Medicine, Virginia Tech, Blacksburg, Virginia 24061, USA David L. Eisenbrandt, Dow AgroSciences, LLC, Indianapolis, Indiana 46268, USA J. Charles Eldridge, Department of Physiology and Pharmacology, Wake Forest University School of Medicine, Winston-Salem, North Carolina 97157, USA

B.B. Gollapudi, Dow Chemical Company, Midland, Michigan 48674, USA Elliot B. Gordon, Elliot Gordon Consulting, LLC, Princeton, New Jersey 08550, USA F. Guerino, Intervet/Schering-Plough Animal Health, Roseland, New Jersey 07068, USA Thomas R. Hanley Jr., Syngenta Crop Protection, Inc., Greensboro, North Carolina 27409, USA Lindsay Hanson, Pest Management Regulatory Agency, Health Canada, Ottawa, Ontario, Canada Paul R. Harp, Lewisville, North Carolina 27023, USA Michael C. Harrass, Rio Tinto Borax, Boron, California 93516, USA Jane E. Harris, Hoffmann-La Roche, Inc. Wayland J. Hayes, Jr., Kansas Life Sciences Innovations, Kansas City, Kansas

(c) 2011 Elsevier Inc. All Rights Reserved.

Contributors

xxv

Thomas Hertner, Syngenta Crop Protection Schwarzwaldallee 215, 4002 Basel, Switzerland

AG,

Dennis R. Klonne, Toxicology & Exposure Assessment Services, Inc., Raleigh, North Carolina 27613, USA

Frederick G. Hess, BASF Corporation, Research Triangle Park, North Carolina 27709, USA

James B. Knaak, Department of Pharmacology and Toxicology, The State University of New York, Buffalo, New York 14260, USA

William F. Heydens, Monsanto Company, St. Louis, Missouri 63167, USA Ernest Hodgson, Department of Environmental and Molecular Toxicology, North Carolina State University, Raleigh, North Carolina 27695, USA

Mike E. Krolski, Bayer CropScience, Stilwell, Kansas 66085, USA Ian C. Lamb, Pioneer Hi-Bred International, Inc., Johnston, Iowa 50131, USA

L. Holden, Sielken & Associates, Inc., Bryan, Texas 77802, USA

Richard L. Lampman, Illinois Natural History Survey, Champaign, Illinois 61820, USA

Jon A. Hotchkiss, The Dow Chemical Company, Midland, Michigan 48674, USA

Mikael Langner, Department of Dermatology, University of California, San Francisco, California 94143, USA

Susan Hurt, Rohm and Haas Company, Philadelphia, Pennsylvania 19106, USA

Jennifer L. Lantz, Bayer CropScience, Research Triangle Park, North Carolina 27709, USA

Sastry Isukapalli, Environmental and Occupational Health Sciences Institute, a joint institute of UMDNJRW Johnson Medical School and Rutgers University, Piscataway, New Jersey 08854, USA

Dominique Lasserre-Bigot, Bayer CropScience, 06560 Sophia Antipolis, France

Seshadri Iyengar, Bayer Monheim, Germany

D-40789

Qing X. Li, University of Hawaii, Honolulu, Hawaii 96822, USA

Poorni Iyer, California Environmental Protection Agency, Office of Environmental Health Hazard Assessment, Sacramento, California 95814, USA

Jing Liu, Oklahoma State University, Stillwater, Oklahoma 74078, USA

CropScience,

Inge M. Jensen, Cheminova A/S, DK-7620 Lemvig, Denmark Russell L. Jones, Bayer CropScience, Research Triangle Park, North Carolina 27709, USA Bernard S. Jortner, Virginia–Maryland Regional College of Veterinary Medicine, Virginia Tech, Blacksburg, Virginia 24061, USA

Edward D. Levin, Duke University Medical Center, Durham, North Carolina 27710, USA

Edward A Lock, School of Pharmacy and Biomolecular Sciences, Liverpool John Moores University, Byrom Street, Liverpool, L3 3AF, UK Marcello Lotti, Università degli Studi di Padova, Instituto di Medicina del Lavoro, Via Facciolati, 71, 35127, Padova, Italy Curt Lunchick, Bayer CropScience, Research Triangle Park, North Carolina 27709, USA

Hideo Kaneko, Sumitomo Chemical Company, Ltd., Osaka, Japan

A. V. Lyubimov, University of Illinois at Chicago, Illinois 60680, USA

Robert J. Kavlock, U. S. Environmental Protection Agency, Research Triangle Park, North Carolina 95814, USA

Howard Maibach, Department of Dermatology, University of California, San Francisco, California 94143, USA

Iain D. Kelly, Bayer CropScience, Research Triangle Park, North Carolina 27709, USA

Lisa E. Maier, University of Michigan, Ann Arbor, Michigan, and Veterans Administration Medical Center, Ann Arbor, Michigan 48105, USA

Michael P. Kenna, U.S. Golf Association, Far Hills, New Jersey 07931, USA Elke Kennepohl, Kennepohl Consulting Young Soo Keum, Seoul National University, Seoul, Korea Jeong-Han Kim, Seoul National University, Seoul, Korea Loreen Kleinschmidt, Environmental Toxicology Department, University of California, Davis, California 95616, USA

Susan Makris, United States Environmental Protection Agency, National Center for Environmental Assessment, Office of Research and Development, Washington, D.C., 20460, USA Mark J. Manning, Rio Tinto Borax, Boron, California 93516, USA Rex E. Marsh, Wildlife, Fish and Conservation Biology, University of California, Davis, California 95616, USA

(c) 2011 Elsevier Inc. All Rights Reserved.

Contributors

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Melanie Marty, Office of Environmental Health Hazard Assessment, California Environmental Protection Agency, Oakland, California 94612, USA

Merle G. Paule, Ph.D., National Center for Toxicological Research, Food and Drug Administration, Jefferson, Arkansas 72079, USA

Ursula May-Hertl, Syngenta Crop Protection AG, Schwarzwaldallee 215, 4002 Basel, Switzerland

Virginie Payraudeau, Bayer CropScience, 06560 Sophia Antipolis, France

Thomas McCurdy, U.S. Environmental Protection Agency, National Exposure Research Laboratory, Research Triangle Park, North Carolina 27711, USA

Erin C. Peck, University of Washington, Seattle, Washington 98195, USA

Tom McKone, Lawrence Berkeley National Laboratories, Berkeley, California 94720, USA Edward C. Meek, Center for Environmental Health Sciences, College of Veterinary Medicine, Mississippi State University, Mississippi State, Mississippi 39762, USA Louise N. Mehler, California Environmental Protection Agency, Sacramento, California 95814, USA Gary J. Mihlan, Bayer CropScience, Research Triangle Park, North Carolina 27709, USA Thomas B. Moore, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California 95814, USA Marsha K. Morgan, U.S. Environmental Protection Agency, Research Triangle Park, North Carolina 27711, USA Ian C. Munro, Cantox Health Sciences International, Mississauga, Ontario, L5N 2X7, Canada Toshio Narahashi, Northwestern University Medical School, Chicago, Illinois 60611, USA Keiichiro Nishimura, Osaka Prefecture University, Osaka 599-8531, Japan Robert J. Novak, University of Alabama, Birmingham, Alabama 35233, USA William J. Ntow, Department of Plant Sciences, University of California, Davis, California 95616, USA Michael A. O’Malley, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento California 95814, USA Frederick W. Oehme, Kansas State University, Manhattan, Kansas 66506, USA Janet Ollinger, Rohm and Haas Company, Philadelphia, Pennsylvania 19106, USA Thomas G. Osimitz, Science Strategies, Charlottesville, Virginia 22902, USA

LLC,

Muhilan D. Pandian, infoscientific.com & risksciences. net, Arlington, Virginia 22201, USA P.P. Parsons, Syngenta

Alain F. Pelfrène, Charbonnières-les-Bains, France Kimberly Pendino, Middlesex County Community College, Edison, New Jersey 08818, USA Barbara J. Petersen, Exponent, Inc. Washington, D.C., 20036, USA Amanda L. Piccirillo, VJP Consulting, Inc., Ashburn, Virginia 20147, USA Vincent J. Piccirillo, VJP Consulting, Inc., Ashburn, Virginia 20147, USA Joachim D. Pleil, U.S. Environmental Protection Agency, Research Triangle Park, North Carolina 27711, USA Kathryn Ponnock, Middlesex County Community College, Edison, New Jersey 08818, USA Carey Pope, Department of Physiological Sciences, Center for Veterinary Health Sciences, Oklahoma State University, Stillwater, Oklahoma 74078, USA Robert H. Poppenga, University of California at Davis, Davis, California 95616, USA Su-wei Qi, University of Hawaii, Honolulu, Hawaii 96822, USA Ruijun Qin, Department of Plant Sciences, University of California, Davis, California 95616, USA Deborah Ramsingh, Pest Management Regulatory Agency, Health Canada, Ottawa, Ontario, Canada Lawrence W. Reiter, U. S. Environmental Protection Agency, Research Triangle Park, North Carolina, 27711, USA Rudy J. Richardson, University of Michigan, Ann Arbor, Michigan 48109, USA Leonard Ritter, School of Environmental Sciences, University of Guelph, Ontario, Canada Jim E. Riviere, Center for Chemical Toxicology Research and Pharmacokinetics, College of Veterinary Medicine, North Carolina State University, Raleigh, North Carolina 27695, USA John H. Ross, infoscientific.com & risksciences.net, Carmichael, California 95608, USA

(c) 2011 Elsevier Inc. All Rights Reserved.

Contributors

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Karl K. Rozman, University of Kansas Medical Center, Kansas City, Kansas 66160, USA

James T. Stevens, Wake Forest University, School of Medicine, Winston-Salem, North Carolina 27157, USA

Andrew L. Rubin, California Environmental Protection Agency, Sacramento, California 95814, USA

Tammy E. Stoker, U.S. Environmental Protection Agency, Research Triangle Park, North Carolina 27711, USA

Michael K. Rust, College of Natural and Agricultural Sciences, University of California, Riverside, California 92507

W.T. Stott, Dow Chemical Company, Midland, Michigan 48674, USA

Luis O. Ruzo, PTRL West, Inc., Hercules, California 94547, USA Terrell P. Salmon, University of California Cooperative Extension—San Diego County, San Diego, California 92078, USA G.K. (Ghona) Sangha, Lanxess Corporation, Pittsburgh, Pennsylvania 15275, USA

Daniel L. Sudakin, Department of Environmental and Molecular Toxicology, Oregon State University, Corvallis, Oregon 97330, USA Chiyozo Takayama, Sumitomo Chemical Company, Takarazuku, Hyogo 665, Japan Emanuela Testai, Istituto Superiore di Sanità, Viale Regina Elena 299, Rome, Italy

James N. Seiber, University of California, Davis, California 95616, USA

Thomas Thongsinthusak, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California 95814, USA

Frank Selman, Dow AgroSciences, Indianapolis, Indiana 46268, USA

Charles Timchalk, Pacific Northwest National Laboratory, Richland, Washington 99352, USA

Larry P. Sheets, Bayer CropScience, Research Triangle Park, North Carolina 27709, USA

Olga A. Timofeeva, Duke University Medical Center, Durham, North Carolina 27710, USA

Linda S. Sheldon, U.S. Environmental Protection Agency, National Exposure Research Laboratory, Research Triangle Park, North Carolina 27711, USA

Abraham J. Tobia, Aventis CropScience, Research Triangle Park, North Carolina 27709, USA

Marilyn Silva, California Environmental Protection Agency, Sacramento, California 95814, USA James W. Simpkins, University of North Texas Health Science Center, Fort Worth, Texas 76107, USA William Slikker, Jr., National Center for Toxicological Research, Jefferson, Arkansas 72079, USA

Rogelio Tornero-Velez, U.S. Environmental Protection Agency, Research Triangle Park, North Carolina 27711, USA Nicolle S. Tulve, U.S. Environmental Protection Agency, National Exposure Research Laboratory, Research Triangle Park, North Carolina 27711, USA Matazaemon Uchida, Nihon Nohyaku Kawachi-Nagano, Osaka 586, Japan

Company,

Paul Slovic, Decision Research, Inc., Eugene, Oregon 97401, USA

István Ujváry, iKem BT, H-1033 Budapest, Hungary

Andrew G. Smith, Medical Research Council Toxicology Unit, University of Leicester, Leicester, LE1 9HN, United Kingdom

Daniel Vallero, U.S. Environmental Protection Agency, National Exposure Research Laboratory, Research Triangle Park, North Carolina 27711, USA

Wayne R. Snodgrass, University of Texas Medical Branch, Galveston, Texas 77555, USA

Bennard van Ravenswaay, BASF AG

Jon R. Sobus, U.S. Environmental Protection Agency, Research Triangle Park, North Carolina 27711, USA David M. Soderlund, Cornell University, Geneva, New York 14456, USA Keith R. Solomon, University of Guelph, Guelph, Ontario, N1G 2W1 Canada Frank Spurlock, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California 95814, USA

Felix Waechter, Syngenta Crop Protection Schwarzwaldellee 215, 4002 Basel, Switzerland

AG,

Edgar Weber, Syngenta Crop Protection Schwarzwaldellee 215, 4002 Basel, Switzerland

AG,

Ronald C. Wester, University of California, San Francisco, California 94143, USA Paul Whatling, Cheminova, Inc., Arlington, Virginia 22209, USA Gary K. Whitmyre, risksciences.com, LLC, Arlington, Virginia 22201, USA

(c) 2011 Elsevier Inc. All Rights Reserved.

Contributors

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Heinrich Wicke, Bayer CropScience, D-40789 Monheim, Germany

S.R. Yates, USDA-ARS United States Salinity Laboratory, Riverside, California 92507, USA

Sanjeeva J. Wijeyesakere, University of Michigan, Ann Arbor, Michigan 48109, USA

Masanori Yoshida, Nihon Nohyaku Company, KawachiNagano, Osaka 586, Japan

Martin F. Wilks, Swiss Centre for Applied Human Toxicology, 4031 Basel, Switzerland

Bruce M. Young, Bayer CropScience, Research Triangle Park, North Carolina 27709, USA

Alan G. E. Wilson, Pharmacia Corporation, St. Louis, Missouri 63167, USA

Frank G. Zalom, Department of Entomology, University of California, Davis, California 95616, USA

Barry W. Wilson, University of California, Davis, California 95616, USA

Valerie Zartarian, U.S. Environmental Protection Agency, National Exposure Research Laboratory, Research Triangle Park, North Carolina 27711, USA

Michael D. Woodward, DuPont Crop Protection, E. I. du Pont de Nemours and Company, Newark, Delaware 19702, USA Jayne Wright, Syngenta Crop Protection, Inc., Bracknell Berkshire RG42 6EY, United Kingdom

Hongbo Zhai, Department of Dermatology, University of California, San Francisco, California 94143, USA Xiaofei Zhang, General Dynamics Information Technology, Henderson, Nevada 89052, USA

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Foreword

Paracelsus—dose response “Alle Ding sind Gift und nichts ohn Gift; alein die Dosis macht das ein Ding kein Gift ist” [all things are poison and not without poison; only the dose makes a thing not a poison”]. With the exception of E  mc2, perhaps no other single statement has wielded such force in establishing the popular notoriety and the professional stature of an individual in the history of science as the words just quoted. In 1993 the New York Academy of Medicine Library exhibited Paracelsus’s works to commemorate the 500th anniversary of his birth. Edward Farber identifies Paracelsus as “the figure head of the 16th century” in The Evolution of Chemistry (Farber, 1952). Reynolds Historical Library (University of Alabama, 1999) curator Marion G. McGuinn writes that “it would be difficult to imagine the healing art as we know it today apart from the historical influence [Paracelsus] brought to bear” (http://www.uab.edu/reynolds/ parcels.html). In his study of Paracelsus’s importance to pharmacology, Mark Young (2004) writes “Paracelsus and his followers caused the pharmacopoeia to be rewritten.” A text on pharmacology in nursing credits Paracelsus with exerting “a profound influence upon the medical beliefs of his time and of succeeding centuries” (Bergersen and Goth, 1979). In Remington’s Pharmaceutical Sciences, Higby (1990) lauds Paracelsus for “sparking the growth of the modern pharmaceutical sciences.” Countless textbooks, handbooks, encyclopedias and dictionaries (general and special), and monographs give him similar credit. Other sources refer to him as the “Father of Toxicology.” Furthermore, his name appears as a significant figure among voluminous numbers of works on homeopathy, natural medicine, alternative medicine, and botanical studies. A web-based Paracelsus mailing list is part of a “health web system,” and the Paracelsus Healthcare Corporation runs the Bledsoe County General Hospital of Pineville, Tennessee. On the popular front, the Stiegl’s Beer (Salzburg, Austria) page offers “Paracelsus Naturtrueb—the unfiltered beer specialty” [(1999) http://www.stiegl.co.@/ebeer.htm]. This is a fitting tribute to Paracelsus’s interest in “naturalness” versus the artificiality he observed in the academic world of his time, but it is also ironic. Behaving in accordance with

his alchemical tendencies, Paracelsus probably would have filtered the beer. The ironies of Paracelsus’s reputation are legion. In fact, it is a wonder that an individual like Paracelsus should enjoy the attention of 21st century readers at all, much less those readers seeking essential information about pesticide toxicology. His contemporaries often found his behavior and theories enigmatic at best. At worst they regarded them as heretical, bizarre, and contentious. Scholars and critics in nearly every century since his death have maligned him, yet others have hailed him as a courageous visionary pioneer equal to Martin Luther. Such conflicting assessments have arisen in part from a variety of misunderstandings of his copious but cryptic writings. Although there exists little disagreement that he was rebellious, iconoclastic, and stubborn, textual evidence supports the view that he made some astoundingly insightful discoveries. Joseph F. Borzelleca (1999) has, perhaps, assessed Paracelsus rightly by calling him the “herald” of modern toxicology, trumpeting his views on “many fundamental issues such as the meaning of life and death, health and the causes of disease (internal imbalances or external forces), the place of humans in the world and in the universe, and the relationship between humans (including himself) and God.” His political heresies and clumsiness toward the medical establishment embittered many of his contemporaries and even some of those who would have been his colleagues. Paracelsus was suspicious and protective of his “special” alchemical knowledge, characteristics that might otherwise have been tolerated, but in Paracelsus they made others respond negatively. It is little wonder that the commentary estimating Paracelsus’s importance and contributions is wide-ranging. In her work describing science in the Renaissance, Marie Boas (1962) attributes to Paracelsus a parental influence on modern chemistry. Bernard Jaffe (1976), who devotes a full chapter of his historical study of chemistry to Paracelsus, claims that the world owes Paracelsus a debt for having planted the seeds of laboratory science. He writes that Paracelsus’s most significant contribution was “not one epoch-making discovery” but

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Foreword

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instead a “vital impetus” which he initiated by “sweeping aside the teachings of the ancient authorities” and “[bringing] alchemy to the aid of medicine.” Many others agree with Jaffe, crediting Paracelsus with the innovation of “iatrochemistry” or “chemistry in the service of medicine” (Partington, 1957). Wolfgang Schneider takes issue with this assessment, specifying that the term “chemiatry” fits the work of Paracelsus: “Oswald Croll (c. 1560–1609) wrote the first textbook of chemiatry, named Basilica chymica which was later revised and enlarged by … Professor Johann Hartmann (1568–1631), instructor in chemiatry at Marburg. In the work of Johann Baptist van Helmont (1577–1644) the influence of Paracelsus is also evident” (Schneider, 1980). It was in the context of chemistry, the study and use of chemicals for medical purposes, that Paracelsus gained particular notoriety, a renown that has continued to the present day. For example, a mid-1990s advertisement for Index Chemicus includes a reference to Paracelsus as an individual who made “legendary discoveries,” specifically the use of chemicals for medicine. By association the advertisement places Paracelsus in the company of Anders Celsius, John Dalton, Adedeo Avogadro, Marie Curie, and T. W. Richards. On the other hand, the head of one prestigious medical school in the U.S. commented in a 1998 interview that “medical schools generally consider his work to be irrelevant.” He added that “those who recognize the name of Paracelsus, … recognize his insight that chemical manipulation of drugs could improve their efficiency” (Reigle, 1998). Much of the controversy connected with Paracelsus has to do with his outspoken challenges to the prevailing practices of physicians, surgeons, and apothecaries. He believed that many of these practitioners sought personal gain above the welfare and safety of the infirm, afflicted, and occupationally exposed masses: “There are two kinds of physician—those who work for love, and those who work for their own profit. They are both known by their works; the true and just physician is known by his love and by his unfailing love for his neighbour” [Selected Writings (Paracelsus, 1951)]. Paracelsus believed that God gave herbs, a term he used in the very broad generic way of the 16th century, “power and virtue to free man from his infirmity” and to “protect his life span against the wrath of death up until the last minute” (Paracelsus, 1951, emphasis added). Thus, Paracelsus espoused the ideas of safety and restoration. He sought to help people avert suffering, to ward off affliction, or if it were not possible to ward it off then to turn it away by curing it once the victim had been afflicted (Paracelsus, 1951). He aimed his studies and his advice toward protection from infirmity, asking rhetorically how one could possibly “protect himself from harm and disaster if he does not know his enemy” (Paracelsus, 1951). Further, if protection and cure failed, at least the physician could prevent the disease from getting worse. To know the enemy, it was essential to study the conditions

in which the afflicted persons lived and worked. This was a bold assertion meant to challenge the prevailing notion that all disease arose from imbalance or overabundance or underabundance of one of the four humors. Paracelsus expressed great concern for the health of all people, high-caste and ordinary: “… to love the sick, each and all of them, more than if my own body were at stake.” This concern seems to have arisen when he worked among the miners of his native Einselden, alongside his father, as Paracelsus sought cures for “miner’s disease” (Noble, 1992). In his works, he admits to being different, claims the superiority of his knowledge and methodology over the ancients, and casts aspersions on the medical profession: “I for my part am ashamed of medicine, considering what an utter fraud it has come to be.” He pursued medical knowledge with the fervor of a zealot: “This is my vow: To perfect my medical art and never to swerve from it so long as God grants me my office, and to oppose all false medicine and teachings” (Paracelsus, 1951). He vigorously opposed the classical teachings and practices. Conventional “humoral” physicians of the 16th century were more concerned with the “accumulation of learning” than with treating disease. They were also more concerned, he asserted, with the unscrupulous accumulation of wealth: Since such useless rabble befoul the art of medicine with their bungling, and seek nothing but their own profit, what can it avail that I admonish them to love? I for my part am ashamed of medicine, considering what an utter fraud it has come to be. (Paracelsus, 1951)

Casarett, Klaassen, and Doull summarize Paracelsus’s principles very well: [Paracelsus] promoted a focus on the ‘toxicon,’ the primary toxic agent, as a chemical entity, as opposed to the Grecian concept of the mixture or blend. A view initiated by Paracelsus that became a lasting contribution held as corollaries that (1) experimentation is essential in the examination of responses to chemicals, (2) one should make a distinction between the therapeutic and toxic properties of chemicals, (3) these properties are sometimes but not always indistinguishable except by dose, and (4) one can ascertain a degree of specificity of chemicals and their therapeutic or toxic effects. These principles led Paracelsus to introduce mercury as the drug of choice for the treatment of syphilis, a practice that survived 300 years but led to his famous trial.

Medieval doctors concentrated on the learning collected in works by Galen, Avicenna, and other classical ancients. Contrary to the accepted canon of his time, Paracelsus, whose influences included Hippocrates, believed that finding the right way to cure an ailment or an infirmity was the physician’s paramount mission: “to love the sick, each and all of them, more than if my own body were at stake” and not “to administer any medication without understanding, nor to collect any money without earning it” (Paracelsus, 1951).

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Foreword

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In the 16th century these were bold, accusatory words, but Paracelsus was a bold man. That boldness—some then and now regard it as ego and audacity—moved him to break with tradition, seeking an effective method and useful knowledge on a combination of trial and error and wideranging testimony from almost any source he could enlist: “… wherever I went I eagerly and diligently investigated and sought after the tested and reliable arts of medicine. I went not only to the doctors, but also to barbers, bathkeepers, learned physicians, women, and magicians who pursue the art of healing” (Paracelsus, 1951). He sought to understand Nature and the nature of its elements and essences. For the wisdom that he believed would result, he traveled widely: Wisdom is a gift from God. Where he gives it, there should one seek it out … . For this I would prove through nature: he who would explore her, must tread her books with his feet. Scripture is explored through its letters; but nature from land to land. Every land is a leaf. Such is the Codex Naturae; thus must her leaves be turned. [Four Treatises (Paracelsus, 1996)]

His contemporaries may not have shared his sense of “experiment”: “Every experiment is like a weapon which must be used in its particular way: a spear to thrust, a club to strike. Experimenting requires a man who knows when to thrust and when to strike, each according to need and fashion” (Paracelsus, 1996). He felt that he learned “when to thrust and when to strike” by experiencing Nature directly and by inquiring of the people he met during his travels: “The book of Nature is that which the physician must read; and to do so he must walk over the leaves.” Moreover, Paracelsus’s ideas of scientia and experientia differed radically from their typical 16th century usages. “In a characteristic reversal of traditional social and intellectual categories, Paracelsus defined scientia as the ‘virtue present in natural objects,’ which the physician must ‘overhear’ and with which he must achieve union” (Smith, 1994). On this basis, he built the principle and raised to a new level of significance the theory that sense observation and experience can—must—provide the proof of the physician’s hypotheses. What he gleaned from his wide travels and conversations, if not seminal, also stands apart from the work of most other medieval scientists by virtue of its being very unusual and pioneering. References to his personality and scientific contributions to modern medicine, modern chemistry, psychiatry, pharmacology, and toxicology have appeared in countless histories of science, science textbooks, scientific handbooks, and articles. According to the website http://www. paracelsian.com/paras.html: He is considered the father of Toxicology. It was Paracelsus’s belief that it was not the substance that was toxic but the amount that was toxic. Paracelsus believed that everything, in excess, was potentially toxic. Conversely, he

believed that certain substances such as arsenic, mercury, lead, etc. could be beneficial in the treatment of disease if administered in very small, controlled dosages.

This quotation contains the most often cited and most widely misunderstood bit of Paracelsean insight. Probably because the insight derived from an intuitive base rather than a modern experimental base, Paracelsus has not received credit for his understanding of dose response. He comprehended, applied, and wrote about hormesis, or the beneficial use of toxic substances in small doses, but he followed some rather bizarre ideas and reasoning to support his contention. On the one hand, he experimented successfully with low-dose applications of mercury to treat epilepsy; on the other hand, he claimed that the source of the epilepsy sprung from a bubble that ascended to the brain and burst there. Arndt-Schulz established an experimental basis for the phenomenon that Paracelsus contrived in his peculiar alchemical way. Medical historians and others have connected his name with a wide range of therapies, procedures, and applications: electroconvulsive therapy (Yudofsky et al., 1991); homeopathy (Ullman, 1988); use of opium and laudanum as painkillers (Levinthal, 1988; Porter, 1996); and aromatherapy (Jacobs, 1996). Some commentaries, including the aforementioned study of pharmacology in nursing (Bergersen and Goth, 1979), credit him with pioneering work as a physician who revolted against the archaic practices of his contemporaries with clumsy, faltering, but remarkable application of dose–response principles (Magner, 1992). Others name him as the most significant individual force in the decline of alchemy and its cryptic mysteries (Barnes-Svarery, 1995). Nonetheless Paracelsus’s writing about “poisons” and “dose” has captivated toxicologists and pharmacologists. The most widely quoted statement about the topic appeared in Paracelsus’s “Third” Defence Concerning the Description of the New Receipts,” published posthumously in 1564. The “Third” is one of seven defenses presented as replies to the accusations of Paracelsus’s enemies. These enemies are the medical establishment of his day, particularly the scholastic physicians and apothecaries. Paracelsus accused them of growing rich at the expense of the poor, of being narrow- and closed-minded, of making faulty diagnoses, and of lacking true knowledge and piety. The establishment, in its turn, questioned Paracelsus’s diagnoses, his itinerancy, his contentiousness, and his “poisonous” prescriptions. C. Lilian Temkin translates the title “Sieben Defensiones, Verantwortung über etliche Verunglimpfungen seiner Mißgönner” as “Seven Defenses, the Reply to Certain Calumniations of His Enemies” (Paracelsus, 1996). Using a common rhetorical strategy of the epoch, Paracelsus refers to himself in the third person in this and other writings. His word choice may warrant a bit more stridency than Professor Temkin has given it in her translation. “Verantwortung,” for example, includes the sense of vin-

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Foreword

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dication and defensiveness, perhaps a vengeful reply. It might be used when matters of honor and reputation are questioned. “Verunglimpfungen” also expresses harsh connotations associated with defamation and disparagement. He aimed his barbs at physicians who “did not practice the right way”: “For what reason is the physician’s calling practised with so much stupidity and so little art, although he fancies himself so important and so superior” [Selected Writings (Paracelsus, 1951)]. In his defense he repeats his diatribe against the apothecaries: Thus too come all the lazy and profligate rascals into medicine and sell their medicine, whether it makes sense or not … . Thus the apothecaries too and some barbers take medicine upon themselves, behave and carry on as though it were a woodcart, go into medicine against their own conscience, forget their own soul, if only they become rich, prepare house and home and all that belong in it, and dress it up. [Four Treatises (Paracelsus, 1996)]

“Calumniations” certainly fits the paranoiac quality one may sense in these defenses, suggesting that Paracelsus felt that he had been tricked or deceived and that his enemies were tricking and deceiving the public as well as his colleagues. By using “Mißgönner,” Paracelsus accusingly implies his feelings of envy and grudging recognition toward his enemies. Paracelsus’s appointment as professor in the University of Basel exemplifies the contentious nature of his career. The position began as a political reward for Paracelsus’s treatment of John Froben’s infected leg. Paracelsus recommended against amputation and saved the leg and the man, witnessed by Erasmus, who was a friend of Froben. Led by Erasmus, Froben’s associates secured the university position for Paracelsus, but many of the Basel faculty were not present to hear and consent to the appointment. Thus Paracelsus arrived somewhat “uninvited.” Further exacerbating the circumstances, Paracelsus began his tenure by immediately announcing his disfavor of the “scholastic” approach to medicine, his disdain for Avicenna and Galen (demigods of that epoch’s medical world), and his dislike for other traditions such as lecturing in Latin and closing lectures to barber-surgeons. A course of events that included Paracelsus publicly insulting a judge—a potential capital offense—led to Paracelsus’s flight from Basel after proving himself to be litigious and caustic to friends and enemies alike. All were in some way or another out to get him, he thought. All seemed to question his diagnoses and his prescriptions. In his second defense, Paracelsus claims “authority” for making his unorthodox prescriptions. His tone and words leave no room for doubt that he sees himself as superior to others, an attitude that did not endear him to his peers: To everyone it is given to speak to advise and to teach, but it is not given to everyone to speak and teach things of strength. For you know that the Gospel too testifies that when Christ taught, He spoke as One who had authority

and not as the scribes and hypocrites. Such authority one should respect as proves itself with works, if one is incredulous of the word. [Four Treatises (Paracelsus, 1996, p. 17)]

The question of authority was a very important matter in Paracelsus’s era. Luther’s reform ideas included the notion of ad fontes (“back to the sources”), in other words going back to the biblical text rather than to the authority of the Pope and the Church and its clerics for knowledge and understanding and interpretation of the spiritual world. Similarly, the concept included going back to the Greek philosophers rather than relying on moderns for knowledge and understanding of the material world. So important and so widely accepted was this notion that such ancients as Avicenna and Galen were revered; memorizing and repeating their ideas and practices were regarded as the highest forms of medical ability. During the 16th century an astonishing 590 editions of Galenic treatises appeared, the main publishing centers being Paris, Lyons, Venice and Basel (Porter, 1997). Paracelsus had been attacked for lack of respect for the medical establishment, for his slow diagnoses, for his itinerant habits, and for his lack of civility. Apothecaries despised him for questioning and exposing their money-mongering practices. Other physicians questioned his use of information drawn from ordinary people (midwives, shopkeepers and butchers, simple country-folk, and mine workers). Scholars rebuked him for his heresies, as exemplified by his burning books of the ancients, in particular Galen. While it is true that Paracelsus paid attention to a variety of factors in treating individuals and that one of his considerations involved “dose,” modern readers must be wary of equating his defensive comments with an exposition on dose–response protocols. For Paracelsus the idea of dose may be seen as theological rather than chemical: Man consists of the four elements, not only—as some hold—because he has four tempers, but also because he partakes of the nature, essence, and properties of these elements. In him there lies the “young heaven,” that is to say, all the planets are part of man’s structure and they are the children of the “great heaven” which is their father. [Das Buch Paragranum (Paracelsus, 1996)]

Alchemists held this macrocosmic–microcosmic view of life quite commonly; for many of them their chief goal was to bring these two worlds into closer connection, in fact to restore the perfect impression the macrocosm had on the microcosm. The latter was believed to be the image and likeness of the former, after all. Further, he thought that all of creation emanates from God, even what appears to be negative or destructive from the human perspective: Is not a mystery of nature concealed even in poison? … What has God created that He did not bless with some great gift for the benefit of man? Why then should poison be rejected and despised, if we consider not the poison but its curative virtue? … And who has composed the prescriptions of nature? Was it not God?

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It is in this sense and context that Paracelsus made the statement that has reverberated his reputation. For it was this topic more than any other accusation that generated the most controversy related to Paracelsus’s new prescriptions. His enemies alleged that Paracelsus used poison to treat his patients: “the use of inorganic, particularly metallic, elements in internal remedies was attacked as unnatural and poisonous …” (Paracelsus, 1996, p. 6). It is this allegation and the quarrels over this use of metals that provided the impetus for Paracelsus’s “third” defense. He opens that defense with a characteristic portrayal of the establishment as being composed of men of little understanding and knowledge of natural forces (Paracelsus, 1996, p. 21): If you wish justly to explain each poison, what is there that is not poison? All things are poison, and nothing is without poison: the Dosis alone makes a thing not poison [emphasis added]. ‘Alle Ding sind Gift und nichts ohn Gift; alein die Dosis macht das ein Ding kein Gift ist.’

First, this quotation has received considerable attention and attained remarkable status among members of the scientific community. It appears here in its original German form with a fairly literal translation. Among its noteworthy features is the fact that it uses the word “Ding,” which means “thing;” it does not use element, substance, medicine. The vague generality of Paracelsus’s usage in this instance seems significant because Paracelsus was an alchemist with the alchemist’s typical mix of mystical and experiential ideas. Like other alchemists, Paracelsus believed that the “Great Work” involved both material and spiritual elements. “Alle Ding” includes the entire extent of “prima materia”: “The still undifferentiated primal substance … everything that is in or has returned to its original state, as well as the unconscious initial state of the soul before it has attained fulfillment, that is to say, ‘before the removal of its dross’” (Paracelsus, 1951). All things are poison—yes; but all things contain an essence, the active good principle in them, often designated as their virtue or power known as their “essentia,” or essence. This “essentia” represented the opposing quality to their “poison.” As he wrote elsewhere [Die drei Bücher des Opus Paramirum (Paracelsus, 1951)], “There where diseases arise, there also can one find the roots of health. For health must grow from the same root as disease, and whither health goes, thither also disease must go.” Paracelsus also maintained the belief that several influences or entia governed humans’ bodies and could do violence to them: ens astorum, the influence of the stars; ens naturale, the natural constitution; and ens veneni, the influence of poison. Also significant is the fact that the quotation emphasizes the inherent toxicity of all things. Most modern readers have inferred that Paracelsus meant something chemical. While it is true that Paracelsus recognized that “dosage” as quantity made a difference in efficacy, he comprehended this as a matter of balance or compensation.

His medicine, his theology, involved the idea that “Each natural disease bears its own remedy within itself” and the principle that “Man has received from nature both the destroyer of health and the preserver of health.” Because God created the whole system, Paracelsus reasoned, it is a perfect system. It may fluctuate from its perfect balance, and when it does humans should try to restore it by searching the resources God has provided, that is, nature. Nature, in turn, also contains both good and bad; it is incumbent upon the physician to use an appropriate measure of Nature’s resource to achieve that stasis: There is always some remedy, a herb against one, a stone against another, a mineral against one, a poison against another, a metal against one, something else against another.

Just as everything (“alle Ding”) contains its essence, everything contains its entia. Paracelsus experimented with dose; he studied degrees of exposure. He considered disease an organic quality rather than a matter of humoral balances. In writing about miners’ sickness, for example, he argued that “the sickness” arose in different ways depending on each miner’s degree of exposure and the kind of mining he did: “… those who work with iron succumb to the spirit of iron and those who work in fire with copper succumb to the copper spirit” (Paracelsus, 1996). It may seem very odd to modern eyes to see Paracelsus refer to the “copper spirit” and the “spirit of iron.” These references appear quaint and perhaps superstitious. In addition to this sense of “spirit,” Paracelsus espoused the notion that all substances consisted of the four classic Greek elements—earth, air, fire, and water—as well as the three Arabic qualities—mercury, sulfur, and salt. Modern readers often misunderstand the former group by thinking of them as the substances they name. In fact they refer to the “essences” or inherent qualities of, for example, fire: volatility, heat, light, consumptiveness, etc. This explains why redheads have suffered the assumption of their having volatile tempers. To many alchemists the association of “red” with fire and a “hot temper” would have been an easy equation. This also explains why “eye of newt” appears in alchemical concoctions: the newt often being a ruddy-skinned creature could lend its “fiery” quality to a mixture. The three qualities—mercury, sulfur, and salt—are common stumbling blocks of Paracelsus’s alchemical writings because in his writing they refer to qualities or “principles” as well as ordinary substances. Sulfur expressed the “spirit of gold” by reason of its color, and the “spirit of fire” because of its combustibility. In this respect, Paracelsus aligned with other alchemists: “… the fundamental thesis so different from our own is the conception that the essential and technically important thing in metals is not material but spiritual” (Hopkins, 1967). Although it may comfort us to know that Paracelsus held views similar to those of many of his contemporaries, the theories of alchemy are such a chaotic mass of ideas derived

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from so many sources and projected onto such a confused background of late medieval adaptations derived from classical and other works that compiling a clear set of systematic rules from them presents insurmountable problems. Furthermore, alchemists typically wrote their formulas and other notes with deliberate ambiguity. They rationalized this latter practice on the grounds that clear notes might provide an unscrupulous or careless practitioner the means to diabolically or inadvertently wreak havoc. In spite of all the quaint arcanities, Paracelsus holds significance for us today as a representative figure. He stands as an emblem of the churning dynamics of his own times, while he inspires modern readers to pursue their investigations with passion and determination. Although the exact date of Paracelsus’s birth remains uncertain (probably November 10, 1493), his epoch unquestionably evokes qualities of momentous proportions. Coincident with Paracelsus’s birth year, Columbus had returned from his first trans-Atlantic voyage and had already set sail for a second adventure. Leonardo DaVinci and Sandro Botticelli flourished as artists. The Treaty of Tordesillas, June 7, 1494, would provide impetus for European exploration, exploitation, and colonization of unparalleled scope and impact. Printing technology and applications had created the possibility of mass communication and wide exchanges of ideas, as well as shifts in the languages used for commerce and politics. Exploration, travel, and communication reached stunning levels by comparison with the levels of these activities in earlier eras. The social climate varied considerably from town to town, but some qualities of social interaction, fashion, and civic interchange give us a sense of the common feelings of the time. The rapid increase in the number and frequency of markets suggests that the atmosphere was dynamic. “All news, political or otherwise, was passed on in the market. In 1534, the actions and intentions of Henry VIII were criticized aloud in the marketplace of Fakenham in Norfolk” (Braudel, 1979, II). Towns tried vigorously to control markets by regulating prices, participants, market days, and locations, but the demand for market exchanges of goods as well as the desire for social dialogue overwhelmed most of their efforts. Sixteenth century European towns were usually defined by protective walls, sometimes moats too. Entrance by strangers required permission from authorities. Yet those in power could exert only so much power against the inevitable: “Traditional habits and customs were lost or smashed. Who would have thought that the belly of London or the belly of Paris would cause a revolution? Yet they did so simply by growing” (Braudel, 1979, II). Fashion trends indicate great fluctuations in the social milieu also, resulting in social classes blurring and blending. The political and theological climate felt similar countercurrents through the writings and public acts of such distinctive characters as Martin Luther, Thomas More, Machiavelli, Loyola, and Copernicus. Although Martin Luther’s translation of the Bible into German marks a

notable achievement, it is equally important to note that he did not have to stand alone. He completed his New Testament in September 1522, his complete Old and New Testaments in 1534. To accomplish this, he worked “in collaboration with a committee of colleagues” (Bainton, 1950). Luther’s courageous outcry against the weaknesses of the Church are well known. Yet in the midst of such efforts as Luther’s, much tradition maintained its stalwart position and sometimes stultifying influence. For example, Latin continued as the expected language of most scholarly intercourse and much internal politics. Maximilian I seemed proud to illustrate his pleadings with his troops in several languages; he had learned German as a child, Latin in school, Saxon and Czech from his subjects, French from his wife, Mary of Burgundy, and Flemish from the officials in the Netherlands. He added Spanish, Italian, and English to his linguistic repertoire by the necessities of diplomacy and military ventures. By 1526, faculty gave law lectures in London’s Inns of Court in English, and “according to statutes setting up Sir Thomas Gresham’s College in 1596, the Monday lecture on medicine was to be given in Latin in the morning and repeated in English in the afternoon because ‘the greatest part of the auditory is likely to be of such citizens and others as have small knowledge or none at all of the Latin tongue’” (Hall, 1994). That Paracelsus raised the hackles of his colleagues at the University of Basel by lecturing in German probably had more to do with his insistence on wearing the alchemist’s apron rather than the scholar’s gown to address his students than with the scholarly inertia of his day: “For a peculiar collaboration between science and esoteric tradition was in fact the norm of the Renaissance, and played an indispensable role in the birth of modern science” (Tarnas, 1991). For thinkers and practitioners like Philippus Aureolus Bombastus Theophrastus Paracelsus von Hohenheim, the times offered remarkable challenges and opportunities. In accomplishing these ends, Paracelsus exhibited what seems to be a characteristic temperament of the time. He shared a mystical religious outlook with his predecessor Roger Bacon (c. 1220–1292) and hermeticist Giordano Bruno (1548—1600), whose Spaccio de la bestia trionfante forecasted a universal moral and religious reform based upon alchemical principles. Paracelsus’s vision of the parallels between and influences of the macrocosm of the universe and the microcosms of Earth and individual humans reflects his faith in the doctrine of signatures: “There is nothing that nature has not signed in such a way that man may discover its essence … . The same is true of man … . Man is endowed with a form corresponding to his inner nature” (Paracelsus, 1951): Since man is a child of the cosmos, and is himself the microcosm, he must be begotten, each time anew, by his mother.

Besides the Neoplatonic and Pythagorean mathematical mysticism and Sun exaltation that ran through all the

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major Copernican astronomers, one finds Roger Bacon, the pioneer of experimental science whose work was saturated with alchemical and astrological principles. Giordano Bruno, the polymath esotericist, championed an infinite Copernican cosmos. William Gilbert’s theory of the Earth’s magnetism rested on his proof that the world soul was embodied in that magnet. William Harvey believed his discovery of the circulation of the blood revealed the human body to be a microcosmic reflection of the Earth’s circulatory systems and the cosmos’s planetary motions. Descartes found support for his theories in an affiliation with mystical Rosicrucianism. Newton derived support from the Cambridge Platonists, and his belief that he worked within an ancient tradition of secret wisdom dates back to Pythagoras and beyond. Indeed the law of universal gravitation itself is modeled on the sympathies of Hermetic philosophy. The modernity of the Scientific Revolution was in many ways ambiguous (Tarnas, 1991). It was a bold age, and Paracelsus was a bold character. His name, Theophrastus Bombastus Philippus Aureolus Paracelsus von Hohenheim, suggests a great deal about his character and manner. The name “Paracelsus,” like the man, derives from a variety of sources, or so we presume. Though it may mean “greater than Celsus,” its origin and meaning remain mysterious. Because Celsus, a first-century-ad Roman medical author, espoused the principles of health according to the humors, theories which von Hohenheim disdained and disputed, it would not have been out of character for the man or for others of his time to accept “Paracelsus” as an epithet for his superiority. Just as likely, “Paracelsus” may simply be the Greco-Roman transliteration of “Hohenheim”: “higher than the sky” or “beyond the sky.” In this sense, it may come from the common medieval practice of name augmentation. Such “an eke name,” a nickname in modern parlance, would probably have suited Paracelsus’s pride about himself and about his hometown Einselden, which he regarded as heaven (yet another way of translating “Hohenheim”). For the members of the Renaissance scientific establishment, this name could hold double annoyance. First, it would seem presumptuous in its self-aggrandizement; second, it would seem heretical in its devaluation of the ancients as authorities. But “Paracelsus,” which he may or may not have invented and used, is just part of the story. His entire name is very long, probably the result of the gradual accretion of details about his life and contributions while he lived and after his death. About every part one finds a wide spectrum of opinion and speculation. Definitive information is scarcer. “Aureolus,” for example, suggests gold and glory, both common aspirations among alchemists. It also relates to a heavenly halo, which one from high heaven would wear. Just as plausible and appropriate in Paracelsus’s case, it may refer to an individual with blond hair (a halo of hair on one with male pattern baldness). “Theophrastus” translates as “God’s words,” also an aspiration of medieval alchemists who valued the idealistic notion of The Gospel of John: “In the beginning was the Word, and the Word was

with God, and the Word was God.” This parallels the Old Testament Bible concept of creation as presented in Genesis in which words (“Let there be light”) precede substance. Medieval alchemists, Paracelsus among them, perceived their ultimate mission as the restoration of the perfect state resulting from the succession of words uttered and then made manifest by the Judeo-Christian God. These alchemists put great stock in the power of words as means of connecting the spiritual with the material world. Paracelsus may have given himself the appellation “Bombastus,” which stems from his rustic ancestry, being a variation of “Baumast,” a German word that relates to trees and their branches. Paracelsus thought of himself as a man of the people, and “Baumast” not only rings with his Germanic roots but also links him with the “Bauers,” who were builders, cultivators, and so on. However, some of his contemporaries may have applied the name as an expression of their disdaining judgment. The frequently outspoken Paracelsus caused many scandals, including one in which he asserted that “German was just as refined and dignified a language [as Latin]” (Hall, 1994). On the surface, his name seems to indict him as a self-serving, insensitive, money-grubbing hand at the alchemical grindstone. However, the causes inspiring his investigations and his insistence on pursuing what his intuition and experience told him must be true emerged from deeper and unselfish roots: This is why I expect thanks from no one. For my medical teaching will give rise to two parties. The first will befoul them; these are not of a breed to thank God or me, but rather will curse me wherever possible. The others will thrive so well that for sheer joy they will forget to thank me. This is the fate of the scientist … (Paracelsus, 1951).

He was shrewd and insightful, yet strange and fanciful. In his own time, his views made him an anathema to most of his contemporaries. In our time, his “experiments” simply do not stand up to the scientific scrutiny of laboratory replication and examination; his results cannot be “repeated, confirmed, refuted and indexed, independently” of him personally. However, from the world of alchemical arcana, now regarded as largely unscientific, he produced influential scientific insights that have profoundly impacted modern scientific and medical activity. William C. Krieger Great Falls, MT

REFERENCES Bainton, R. H. (1950). “Here I Stand: A Life of Martin Luther.” New American Library, New York. Barnes-Svarery, (1995). “The NY Publish Library Science Desk, Reference.” Macmillan, New York. Bergersen, B. S. and Goth, A. (1979). “Pharmacology in Nursing.” Mosby, St. Louis.

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Boas, M. (1962). “The Scientific Renaissance.” Harper & Row, New York. Borzelleca, J. F. (1999). Paracelsus: herald of modern toxicology. Toxicol. Sci. 83, 57–64. Braudel, F. (1979). The Wheels of Commerce (S. Reynolds, trans.). Harper & Row, New York. Farber, E. (1952). “The Evolution of Chemistry.” The Ronald Press, New York. Hall, T. S. (1994). Ideas of Life and Matter. University of Chicago Press, Chicago. Higby, G. J. (1990). Evolution of pharmacy. In Remington’s Pharmaceutical Sciences 18th edn, pp. 8–19. Philadelphia College of Pharmacy and Science, Easton, PA. Hopkins, A. J. (1967). “Alchemy: Child of Greek Philosophy.” AMS Press, New York. Jacobs, J., ed. (1996). “The Encyclopedia of Alternative Medicine.” Journey Editions, Boston. Jaffe, B. (1976). “Crucibles: The Story of Chemistry,” 4th edn. Dover, New York. Levinthal, C. F. (1988). “Messengers of Paradise: Opiates and the Brain—the Struggle Over Pain, Rage, Uncertainty, and Addiction.” Doubleday, New York. Noble, D. F. (1992). “A World without Women: the Christian Clerical Culture of Western Science.” Oxford University Press, New York. Paracelsus (1951). “Selected Writings” (J. Jacobi, ed. N. Guterman, trans.). Pantheon, New York. Paracelsus (1996). “Four Treatises of Theophrastus von Hohenheim Called Paracelsus” (H. Sigerist et al., trans; H. Sigerist, ed.). Johns Hopkins Press, Baltimore. Partington, J. R. (1957). “A Short History of Chemistry.” Dover, New York. Porter, R., ed. (1996). “The Cambridge Illustrated History of Medicine.” Cambridge University Press, Cambridge, MA. Porter, R. (1997). “The Greatest Benefit to Mankind.” Norton, New York. Reigle, R. (1998). An Annotated Bibliography of Works about Theophrastus von Hohenheim, unpublished manuscript Schneider, W. (A. G. Debus ed.) (1980). Science, Medicine and Society in the Renaissance Vol. 1. Watson Academic, New York, pp. 24–36. Smith, P. H. (1994). “The Business of Alchemy: Science and Culture in the Holy Roman Empire,”. Princeton University Press, Princeton, NJ. Tarnas, R. (1991). “The Passion of the Western Mind.” Harmony, New York. Ullman, (1988). “Homeopathy: Medicine for the 21st Century.” North Atlantic Book, Berkeley. Young, M. (2004). “Paracelsus : The Philosopher’s Stone Made Flesh.” http://www.nzepc.auckland.ac.nz/authors/young/paracelsus.asp Yudofsky, S. C., Hales, R. E., and Ferguson, T. (1991). “What You Need to Know about Psychiatric Drugs.” Ballantine, New York. Further Reading Achterberg, J. (1985). Imagery in Healing: Shamanism and Modern Medicine. Shambhala, Boston. Bayard, T., trans. and ed. (1991). “A Medieval Home Companion: Housekeeping in the Fourteenth Century.” HarperCollins, New York. Beecher, H. K., ed. (1960). “Disease and the Advancement of Basic Science.” Harvard University Press, Cambridge, MA. Braudel, F. (1981). The Structures of Everyday Life (S. Reynolds, trans.). Harper & Row, New York.

Butterfield, H. (1962). The Origins of Modern Science. New York, Collier. Cassarett, L. J., and Doull, J. (1993). “Cassarett and Doull’s Toxicology,” 5th edn. New York. Casarett, L. J., Klaassen, C. D., and Doull, J. (2001). “Casarett and Doull’s Toxicology,” 6th edn. McGraw-Hill, New York. Chase, A. (1982). Magic Shots: A Human And Scientific Account of the Long and Continuing Struggle to Eradicate Infectious Diseases by Vaccination. Morrow, New York. Clendening, L., comp. (1960). “Source Book of Medical History.” Dover, New York. Cosman, M. P. (1996). “Medieval Wordbook,”. Facts On File, New York. A. G., ed.Debus, (1972). Science, Medicine and Society in the Renaissance Vols 1 and 2. Watson Academic, New York. Deichmann, W. B., Henschler, D., Holmstedt, B., and Keil, G. (1986). Review of “what is there that is not poison?” A study of the Third Defense by Paracelsus. Arch. Toxicol. 58, 207–213. Dreisbach, R. H., and Robertson, W. O. (1987). “Handbook of Poisoning,” 12th edn. Appleton & Lange, Norwalk, CT. Federmann, R. (1970). The Royal Art of Alchemy. Chilton, Philadelphia. Goldwater, L. J. (1973). “Mercury: A History of Quicksilver,”. York Press, Baltimore. Hartman, F. (1918). “Paracelsus,”. Theosophical Pub, New York. Julien, R. M. (1988). A Primer of Drug Action 5th edn. Freeman, New York. Kaptchuk, T., and Croucher, M. (1987). The Healing Arts: Exploring the Medical Ways of the World. New York, Summit. Kerr, F. W. L. (1981). The Pain Book. Prentice-Hall, Englewood Cliffs, NJ. Magnusson, M., ed. (1990). “Chambers’ Biographical Dictionary.” Chambers, Edinburgh. McCabe, V. (1997). “Let Like Cure Like: The Definitive Guide to the Healing Power of Homeopathy.” St. Martin’s, New York. Magner, L. N. (1992). “A History of Medicine.” Dekker, New York. Barnet ed.) (S. (1969). “Doctor Faustus”. New American Library, New York. Mason, S. F. (1962). A History of the Sciences. Collier, New York. Nicholl, C. (1997). “Chemical Theatre.” Akadine Press, New York. Ottobom, M. A. (1984). “The Dose Makes the Poison,”. Vincente Books, Berkeley. Pagel, W. (1982). “Paracelsus,” 2nd rev. edn. Karger, Basel. Pearsall, R. (n.d.). “The Alchemists.” Weidenfeld & Nicholson, London. Polunin, M., and Robbins, C. (1992). “The Natural Pharmacy.” Macmillan, New York. Porkert, M., and Ullman, C. (1982). “Chinese Medicine,” (M. Howson, trans.). Morrow, New York. Siraisi, N. G. (1990). Medieval & Early Renaissance Medicine: An Introduction to Knowledge and Practice. University of Chicago Press, Chicago. Timbrell, J. A. (1982). Principles of Biochemical Toxicology. Taylor & Francis, London. Whorton, J. (1998). M.D. Telephone interview, 15 April 1998. Wingate, P. and Wingate, R. (1996). The Penguin Medical Encyclopedia 4th edn. Penguin, London. Yates, F. A. (1964). “Giordano Bruno and the Hermetic Tradition,” (reprinted 1979). University of Chicago Press, Chicago.

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Preface

The Third edition of the Handbook is renamed the Hayes’ Handbook of Pesticide Toxicology and dedicated to the memory of Wayland J. Hayes, Jr., whose major contributions to pesticide science are chronicled, in part, in the Dedication. The cover design includes the whimsical 3-segmented, 6-legged doodle by Hayes used on the First and Second editions of the Handbook. This edition of Hayes’ Handbook of Pesticide Toxicology includes primarily new and revised chapters concerning fundamentals of the past and new insights gained from more recent research in pesticide science. More complete exposition of the concepts which have guided preparation of these volumes is contained in the Preface to the Second edition included herein. Pests and organisms that would devour our residences, personal property and food supply remain ever-present competitors in human environments. In response, pesticides delivered in developed nations with increasing precision and regulation represent a chemical technology that is refined, extensively used and studied in detail. Chemical exposures, particularly those related to the economic class pesticide, are an analytical reality that remains problematic for many persons in spite of overwhelming environmental monitoring which reveals that exposures occur at levels benign to health. The Handbook is expected to contribute to clarification, and even resolution, of some imperfections or limitations in available knowledge.

Numerous experts, more than 200 in all, have contributed their time and expertise to the Third edition. Their contributions are particularly noteworthy and appreciated in continued times of economic uncertainty, emerging. Regulatory priorities, and considerable instability in private and public institutions as priorities and programs take new forms. The authors have provided in-depth review and exposition of the particular topics that are included in this edition. References will allow interested readers to pursue topics of interest. Each of the Associate Editors, including John Doull, Joop van Hemmen (deceased), Ernest Hodgson, Howard Maibach, Lawrence Reiter, Leonard Ritter, John Ross, and William Slikker is acknowledged and thanked for his important and particular contributions to the development and production of the Hayes’ Handbook of Pesticide Toxicology. These volumes represent the tireless dedication and exemplary service of Helen Vega, Administrative Assistant in the Personal Chemical Exposure Program here at Riverside and Editorial Assistant for the Hayes’ Handbook of Pesticide Toxicology. We are both grateful to Kirsten Chrisman, Rebecca Garay, April Graham, and Caroline Jones of Elsevier who effectively moved the author’s copy to text. Robert I. Krieger University of California, Riverside

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Section I

Pesticide Uses

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Chapter 1

Dose and Time Determining, and Other Factors Influencing, Toxicity Karl K. Rozman, John Doull and Wayland J. Hayes, Jr. University of Kansas Medical Center

1.1  Introduction This chapter is intended as an introduction to the toxicity of pesticides and an evaluation of methods for their study. This chapter is not meant as a set of guides for testing a particular compound intended for a particular purpose. Of course, the importance of such guides is recognized. In addition to offering suggestions about where details on technique may be found, an effort is made to identify (a) parameters in need of special study and (b) the variety and limitations of present approaches to such study. Under the circumstances, it has seemed best to organize the dis­ cussion of techniques conceptually and not in the usual operational way according to acute and chronic tests, der­ mal toxicity, and the like. Briefly, the statistical and other methods for studying toxic reactions in intact animals are discussed in this chapter. Methods for studying absorption, distribution, metabolism, storage loss, and excretion are considered in chapters on “Pesticides Disposition” as well as techniques for measuring different kinds of injury and injury in different tissues. Methods measuring exposure and quantitative metabolism in people under various other practical conditions are also discussed there. We (Rozman and Doull) considered this chapter by Dr. Hayes to be the best summary of thus far recognized fun­ damental principles and unresolved problems of toxicology long before we were asked to revise it. We recognized that any attempt to revise a scholarly activity of such high qual­ ity is not without risk of ending up with an inferior product. Therefore, the decision was reached together with the Editor in Chief that the chapter as written by Hayes (1991) will be retained in its original form, but that we would add a discus­ sion after each section and subsection when such an addi­ tion is warranted, to bring in the more recent developments regarding the role of time and kinetics in toxicology. As Hayes points out in this chapter, ­toxicology needs a unifying

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theory. In the following sections, we outline our proposal for such a theory and indicate at the proper passages how it would amplify Hayes’ writings.

1.1.1  Dose and Time as Fundamental Variables of Toxicity Toxicity (T) is a function of exposure (E), and E is a func­ tion of dose (c) and time (t) [T  f(E(c, t)]. Toxicity is the manifestation of an interaction between molecules con­ stituting some form of life and molecules of exogenous chemicals or physical insults. Consequences of molecular interactions or physical insults may propagate, through cau­ sality chains, all the way to the organismic level (Rozman and Doull, 2000; Rozman et al., 2006). There are two fun­ damental ways to view this interaction: (1) What does an organism do to a chemical? (2) What does a chemical do to an organism? Dealing with the first question led to the development of the discipline of pharmacokinetics, which was later incorporated into some toxicity studies and there­ fore, in that context, it would be more appropriately called toxicokinetics (K). The other question was addressed by the discipline of pharmacology in the form of pharmaco­ dynamic experiments, which again in the context of toxi­ city would be more properly termed toxicodynamics (D). We recognize that the use of the prefixes pharmaco- and toxico- in the context of kinetics and dynamics is problem­ atic because both involve value judgements not compatible with the unbiased interpretation of (a) law(s) of nature. However, before this issue can be sorted out in terms of epistemology, we will be using these terms interchangably and often in the traditional (perhaps incorrect) way. Thus, toxicity (T) may be defined as a function of E, K, and D.

T  f (E , K , D)



Hayes’ Handbook of Pesticide Toxicology



This functional relationship can be described mathemati­ cally by a simple differential equation using the chain expansion:



dT dT dD dK    dE dD dK dE

A definition of toxicity according to Rozman and Doull (1998) runs as follows: “[toxicity] is the accumulation of injury over short or long periods of time, which renders an organism incapable of functioning within the limits of adaptation or other forms of recovery.” This definition implies that toxicity is a function of time in addition to the dose. This concept was already recognized by Paracelsus 500 years ago. A closer scrutiny of the definition of toxi­city indicates that the relationship between toxicity, dose (c), and time (t) is a complex one because toxicokinetics itself is dose- and time-dependent [K  f(c, t)] as is toxico­ dynamics [D  f(c, t)]. It should be noted that the various time-dependencies seldom run on the same timescale. Conceptually, K may also be viewed as a function of the dynamic change between absorption (Abs) and elimi­ nation (El),

K  f (Abs, El)

because it is the ratio between entry rate (absorption) and exit rate (elimination) that determines the time course of a compound in an organism. In the simplest case of an intra­ venous bolus injection (instantaneous absorption), the time course is determined by the rate of elimination alone for a compound obeying a one-compartment model. Usually absorption is faster than elimination, making processes related to elimination (distribution, biotransformation, excretion) rate-determining or -limiting in most instances. In analogy, D may be viewed as a function of the dynamic change between injury (I) and recovery (R),

D  f (I , R )

because it is the ratio of injury to recovery that determines the time course of an adverse effect in an organism. The simplest case for such an injury would be when an organ­ ism would recover from an acute injury in accordance with a one-compartment toxicodynamic model. Again, processes related to recovery are usually slower than the rate of injury. Therefore, more often recovery (adaptation, repair, reversibility) will be rate-determining or -limiting. Most often compounds do not behave in an organism according to a one-compartment model. The reason for this is that elimination from the systemic circulation itself can be a function of excretion (Ex), distribution (Dist) and biotransformation (Bio).

El  f (Ex, Dist, Bio)

When any or all of these processes become rate-limiting, two or multi-compartmental models are needed. Again, in analogy to K, recovery (R) in a D model may not be a simple function of, for example, reversibility (Rv), but could also require repair (Rp). In addition, adaptation (Adp) may also be occurring:

R  f (Rv, Rp, Adp)

In such instances, two- or multi-compartment D analy­ ses are needed to describe the toxicity of a compound that affects any or all of these processes. Absorption and injury can be thought of as being analogous manifestations of K and D. Absorption is a function of site (S) and mecha­ nism (M), as is injury: Abs  f (S , M )

I  f (S , M )



This analysis can be continued all the way to the molecular level. It is clear that any rate-determining step or rate-­limiting steps, originating at the level of molecular interactions, will then propagate through causality chain(s) to the levels depicted in Fig. 1.1, which represents a schematic illustration of this concept. Each of the processes depicted in Fig. 1.1 may be doseand time-dependent although past experiments often failed to demonstrate this, because they were conducted with pre­ ponderant emphasis on one or the other; for example, D was mainly studied as a function of dose and K mostly as a function of time. Time has always been an important factor in design­ ing toxicological experiments, yet time as an explicit vari­ able of toxicity has been afforded very little attention. It is interesting that, after Warren (1900) was severely criticized by Ostwald and Dernoscheck (1910) for his analogy of c  t  k to P  V  k of ideal gases, the entire issue was forgotten. Even though c  t  k kept surfacing repeatedly (e.g., Druckrey and Küpfmüller, 1948; Flury and Wirth, 1934; Littlefield et al., 1980; Peto et al., 1991) an analogy to thermodynamics was not contemplated again, at least not to our knowledge! When we “rediscovered” the c  t  k con­ cept in still another context (delayed acute oral toxicity) this required some reevaluation of the role of time in toxicology in both a historical context and as an independent variable. Ostwald and Dernoscheck (1910)’s analogy of toxicity to an adsorption isotherm is problematic, because adsorption entails processes that are far from ideal conditions. Much more reasonable is Warren (1900)’s analogy to P  V  k for ideal gases as a comparison for ideal conditions in toxi­ cology. Reducing the volume of a chamber containing a given number of molecules or atoms of an ideal gas will decrease the time for any given molecule or atom to collide with the wall of the chamber. This leads to increased pres­ sure, which is simply an attribute of the increased ­ number

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity



Decision Tree Host/Agent Interaction [Toxicity = f (D, K,E)] dT dT = dD dE

I

What does agent do to organism? T D D = f (c, t) K E c t I R Abs R EI Adp Rp Rv Dist Adp Rs Rv Bio Ex

dD dK

dK dE

E = f (c, t) Toxicity Dynamic Process (es) Kinetic Process (es) Exposure Dose Time Injury Recovery Absorption Elimination Adaptation Repair Reversibility Distribution Biotransformation Excretion

What does organism do to agent? K = f (c, t)

Abs

EI

Dist

Bio

Ex

Figure 1.1  Conceptual outline of the decision tree approach.

of molecules per unit volume, which is concentration. Thus c  t  k and P  V  k are compatible with each other if looked at mechanistically. Of course, Ostwald and Dernoscheck’s comparison of toxicity to an adsorption iso­ therm is much closer to the real-life situation of toxicology, where the most frequent finding is that (c  tx  k). These thought experiments and some discussions led to the recognition that toxicologists did everything the opposite way of thermodynamicists. Instead of starting out with the simplest model (ideal gas in thermodynamics cor­ responds to ideal conditions in toxicology experiments) and building into it step by step the increasing complex­ ity of the real world, toxicologists tried to predict from one complex situation to another complex situation. In addition, time as an explicit variable was largely ignored although it is one of two fundamental variables of toxic­ ity (Rozman, 1998). It is unlikely that a better understand­ ing of biological processes at the molecular level alone will lead to improved risk predictions in toxicology, as long as the experimental designs of toxicological stud­ ies provide the wrong reference points for departure from ideal to real conditions. For example, the standard inhala­ tion toxicity protocols (6 h/day, 5 days/week) cannot yield c  t  k because after 6 h of intoxication, there are up to 18 h of recovery, and on weekends there are up to 66 h of recovery, at least for compounds of short half-life. This would require at least two additional functions to correct for departure from kinetic steady state. The real-life situa­ tion is even more complex where departures from the ideal condition (steady state) are highly irregular. Nevertheless, it is reasonable to expect that risk predictions will be possi­ ble for even the most irregular exposure scenarios once the

reference points are established as dose-and time-responses under ideal conditions (toxicodynamic or toxicokinetic). In 25 years of studying the toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and related compounds, the con­ cept of c  t  k did not emerge in any other experimental context except in the two recent subchronic-chronic studies, which were conducted under conditions of toxicokinetic steady state (Rozman et al., 1996; Viluksela et al., 1997, 1998). Nevertheless, a general interest in the role of time in toxicology pervaded the herein-presented line of thinking for many years (Rozman et al., 1993; Rozman et al., 1996; Rozman and Doull, 1998; Rozman, 1998). Most toxicolo­ gists are familiar with Haber’s rule of inhalation toxicology and its applicability to war gases and some solvents. Much less attention has been given to Druckrey’s work, which extended the c  t concept to lifetime cancer studies by oral rather than inhalation exposure. Finally, there is very little cross-referencing of the c  t  k data that were generated by entomologists (e.g., Peters and Ganter, 1935; Busvine, 1938; Bliss, 1940) and those established by toxicologists. History demonstrates that a fundamental relationship in sci­ ence keeps reappearing in different contexts as is the case with c  t  k. During this period many apparent excep­ tions seem to be occurring with no satisfactory explanation. Attempts at generalization usually fail until a commonality is detected among all experiments as in this case among those that yielded c  t  k. This commonality is toxico­kinetic steady state and/or irreversibility of an effect, which of course can be interrelated. Anesthesia, like intravenous infu­ sion, leads to rapid and sustained steady state for compounds of short half-life. Most anesthetics and solvents do have short half-lives and many obey Haber’s rule, except when

Hayes’ Handbook of Pesticide Toxicology



­ easurements are taken while an adaptive process is under­ m way, that is, induction of a protein. Druckrey and the ED 01 study used feeding as a route of exposure, which yields a ­better steady state for compounds of intermediate half-life than, for example, gavage. However, the exponent x in the term of Druckrey’s general formula increases above 1 rap­ idly as the half-life of compounds becomes shorter, because there is intermittent recovery between bouts of feeding. Most of the entomology studies were related to fumigation, which often but not always resulted in fairly rapid steady state. Finally 1,2,3,4,6,7,8-heptachlorodibenzo-p-dioxin (HpCDD), which has a half-life of 314 days (Viluksela et al., 1997) in female rats, yields virtual steady state for a 70-day observa­ tion period after any route of administration but TCDD, with a half-life of 20 days, does not. However, when TCDD’s toxicity was studied under steady state conditions, its sub­ chronic-chronic toxicity also occurred according to c  t  k (Rozman et al., 1993; Saghir et al., 2005).

1.1.2  Definition of Dose and Time Before analyzing dose-time relationships further, it is useful to establish clear definitions of these fundamental variables of toxicity. Historically, neither dose nor time has been defined with clarity as a variable of both toxico­ kinetics and toxicodynamics. It is customary to use the terms acute dose and acute effect as if the two were inter­ changeable. In fact, an acute dose can lead to chronic effects (Druckrey et al., 1964) and multiple doses can trig­ ger a fulminant episode (Garrettson, 1983) of toxicity. In risk (safety) assessment it is always the total dose delivered that is of concern, although in therapeutics the daily dose is often referred to simply as the dose. Therefore, a useful definition of dose in toxicology, would be: n

Dose(c) 

∑ dose rates. n1



According to this definition a single acute dose would rep­ resent the limiting case when the dose rate equals the dose. This definition would be valid for any kind of irregularity in the dosing regimens and is analogous to the definition of dose in radiation biology. Ever since the dawn of human consciousness, mankind has struggled with the notion of time. It is not possible to predict what influence the concept of toxicological time will have on our perception of time. Suffice it to say at this junc­ tion, it is not possible to think of toxicity without the implicit presence of time as a variable, although in toxicity studies, time received only semiquantitative designations (acute, sub­ acute, subchronic, chronic). In fact, one could view organ­ isms as instruments exquisitely sensitive to time. Important for toxicology is the concept that the time course of a toxi­ cant in an organism (kinetics) is very often different than the

time course of toxicity (dynamics). Underlying biological processes (absorption, distribution, elimination, injury, adap­ tation, recovery) have their own timescales depending on the molecular events behind each process (e.g., enzyme induc­ tion, receptor regulation either directly or via gene expres­ sion). Thus, in toxicology the dose is a pure variable, but there are many different processes occurring on different time­ scales yielding different c dt integrals leading to complex interactions, which can be described as c  tx. In spite of this complexity, science can deal with it in a traditional, analyti­ cal fashion. Because only knowledge of rate-limiting steps is required to accurately describe toxicity, this will often reduce complexity to manageable proportions.

1.1.3  Dose and Time Relationships Toxicity is a function of exposure and exposure is a func­ tion of dose and time [T  f[E(c, t)]]. Consequences of interactions between a toxic agent and an organism at the molecular level propagate through toxicodynamic or toxicokinetic–­toxicodynamic causality chains all the way to the manifestation of toxicity at the organismic level (Fig. 1.1). If the recovery (consisting of adaptation, repair, and reversibil­ ity) half-life of an organism is longer than the half-life of the causative agent in the organism then toxicodynamics becomes rate-determining (one-compartment model) or rate-limiting (multicompartment models) (Rozman and Doull, 2001a). If the toxicokinetic half-life of the compound is longer than the recovery half-life, then toxicokinetics will be rate-­determining (-limiting), in which case the toxicokinetic area under the curve (AUC) will be identical to the toxico­dynamic AUC. There are three limiting conditions for c  t  k to emerge when the causality chain propagates through either toxico­ dynamic or toxicokinetic-toxicodynamic processes: Toxicodynamics 1. In case of no recovery (no reversibility, no repair, no adaptation) linear accumulation of injury will occur according to a triangular geometry (c  t/2  k) foll­ owing repeated doses or according to a rectangular geometry after a single dose (c  t  k), provided that the c  t lifetime threshold has been exceeded, which occurs when cthreshold  tlifespan  k. 2. After recovery (reversibility, repair, adaptation) steady state has been reached, injury will occur according to a rectangular geometry (c  t  k), after exceeding the c  t lifetime threshold. Toxicokinetics 1. No elimination will lead to linear accumulation of a com­ pound and as a consequence to accumulation of injury according to a triangular geometry (c  t/2  k) after repeated doses or according to rectangular geometry after a single dose (c  t  k) above the c  t lifetime threshold.

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

2. After toxicokinetic (and as a consequence toxico­ dynamic) steady state has been reached, injury will occur above the c  t lifetime threshold according to a rectangular geometry (c  t  k). Exposure Frequency As the toxicokinetic and toxicodynamic half-lives become shorter and shorter the distinction between elimination and recovery half-lives becomes less important, because another time dependence, that of the frequency of expo­ sure, starts dominating the time dependence: 1. Compounds having very short toxicokinetic or toxico­ dynamic half-lives will reach steady state rapidly and yield c  t  k upon continuous exposure according to a rectan­ gular geometry above the c  t lifetime threshold provided that adaptation and repair are also at steady state. 2. Other types of geometries certainly can be created by elaborate, but regular dosing regimens. These scenarios are less likely to play a practical role in toxi­cology, although they may be of theoretical interest in the establishment of model parameters for predicting toxi­ city after irregular dosing regimens. It should be kept in mind that the mathematics of firstorder processes, when appropriate, are valid for bimolec­ ular reactions (e.g., receptor binding), which result in the propagation of the causality chain to the level of model­ ing (Fig. 1.1). Therefore, 90% of toxicodynamic steady state will not be reached until 3.32 recovery half-lives have elapsed. Thus, Haber’s rule will be obeyed only if the observation period is outside of about 4 recovery half-lives or if recovery is a zero order process. Thus, the various (c  t  k) scenarios represent lim­ iting conditions (Rozman and Doull, 2001b). The magni­ tude of the c  t product is a function of the potency of the compound, of the susceptibility of the organism, and of the deviation from the ideal conditions and will yield c  tx  k for nonlimiting conditions. It should be recog­ nized that the dose (c) does not have inherent exponential properties, but time (t) does have such properties, because under nonideal conditions toxicity is a function of at least two independent timescales: one being the half-life of the rate-determining step (toxicodynamic and/or toxicokinetictoxicodynamic) of the intoxication (intrinsic property of organism or compound), the other one being the frequency (which includes duration) of exposure, which is indepen­ dent of both the compound and the organism. In conclusion, these data and consideration of a signifi­ cant body of evidence accumulated over the last 100 years suggest that c  t  k is part of a fundamental law of toxi­ cology, and possibly of biology in general, that can be seen only under ideal conditions (Rozman and Doull, 2001b). It must be emphasized that the dimension of the c  t prod­ uct is not energy but effect (Rozman, 2008) as in action



(Wirkung). This has been confirmed using other classes of compounds and the herein-described ideal conditions (Saghir et al., 2005). Therefore, Paracelsus’ famous state­ ment should be supplemented to read “Dosis et tempus fiunt (faciunt) venenum” (Dose and time together make the poison). Implications for risk assessment are that the mar­ gin of exposure (MOE) must be defined in terms of both dose and time. This can be done by relating the real-life (discontinuous) exposure scenario to that of ideal (continu­ ous) exposure condition:



MOE 

c  tx ct

The margin of safety and its reciprocal, the margin of risk, can be determined when the MOE exceeds the c  t life­ time threshold (Rozman and Doull, 1999, 2000). Figure 1.1 may also be used as a decision tree to iden­ tify critical steps needed for modeling to predict toxicity. It is important to note that both a high degree of irreversibil­ ity and toxicodynamic steady state are rare phenomena in toxicology, although both can be seen any time the obser­ vation period is much shorter than the recovery half-life. In real-life situations there are usually at least two or three rate-limiting steps in toxicokinetics and likely as many in toxicodynamics. It must be emphasized, though, that mul­ tiple toxicokinetic compartmental models do not necessar­ ily require multiple toxicodynamic models, and vice versa. However, if there are three different rate-limiting processes occurring on different timescales in toxicokinetics and three different rate-limiting processes taking place on three dif­ ferent timescales in toxicodynamics, such a scenario would represent a formidable computational task for a theoretical treatise on risk assessment. Therefore, a practical approach would be to conduct experiments at toxicodynamic steady state (which of course would require a preexisting toxico­ kinetic steady state in many instances) as a point of refer­ ence clearly defined by c  t  k. Then, experiments would need to be carried out for different compounds with differ­ ent half-lives to establish model parameters, which describe departures from toxicokinetic-toxicodynamic steady state of increasing frequency and irregularity. In summary, c  t  k represents the most efficient (a kind of worst case) exposure scenario for producing an effect, namely, continuous exposure till manifestation of an effect. Experimentally, this condition is often met by continuous inhalation exposure or daily oral administra­ tion of compounds that have toxicodynamic-toxicokinetic half-lives of a few days or longer and /or effects that are largely irreversible. It must be emphasized that any depar­ ture from the worst case scenario will result in a change of c  t  k into c  tx  k. Departures are represented by regular or irregular interruptions of exposure and/or inter­ mittent recovery from injury. The larger the departure, the larger will be x, indicating that increasing x is equivalent to

Hayes’ Handbook of Pesticide Toxicology



decreasing toxicity. This is entirely logical, when recogniz­ ing that increasing interruptions of exposure and/or injury will result in longer and longer periods of time needed to cause toxicity equivalent to that of continuous exposure, because of increasing intermittent recovery. To express this more clearly, we can write

c  t x  k or c  t  t x1  k

Thus, tx1 is a simple transforming factor which changes the slope of the log c vs log t plot back to unity. It may be viewed as the toxicological timescale of recovery which runs counter to the timescale of toxicity, thereby reducing it. A limiting condition for first-order processes will be reached when exposure occurs outside of 6.64 toxico­ kinetic-toxicodynamic half-lives, because at that time 99% elimination/recovery will have occurred. Under such con­ ditions (which are closest to the real-life situation for most compounds), toxicity will be less dose- and time-­dependent. In this case mainly the frequency of exposure will deter­ mine x. If x, is then determined experimentally, for say 1, 2, 4, 8, 16, and 32 days for a compound with a toxicoki­ netic or toxicodynamic half-life 3.6 h after continuous vs intermittent exposure under isoeffective conditions, then plotting of the data will allow extrapolation to any exposure scenario outside of 6.64 half-lives (which corresponds to 1 day). Most dietary constituents fall in this category. For zero-order processes two half-lives are needed for elimi­ nation and/or recovery. It should be kept in mind that the half-life of zero order processes (unlike that of first-order processes) is concentration-dependent. A series of articles has explored how other disciplines deal with complex systems (Goldenfeld and Kadanoff, 1999; Koch and Laurent, 1999; Weng et al., 1999; Whitesiles and Ismagilov, 1999). Goldenfeld and Kadanoff (1999) made some important observations which are relevant for toxi­ cology. Simple laws of physics give rise to enormous com­ plexity when the number of actors is very large. We have the same paradox in toxicology in that the c  t concept is very simple, but the “real-world” manifestation of toxicity is very complicated. One other observation is equally relevant: “Use the right level of description to catch the phenomena of interest. Don’t model bulldozers with quarks.” This trans­ lates in toxicology to: Don’t model toxicity at the molecular level. The decision tree approach in Section 1.1.1 (Fig. 1.1) was developed to aid toxicologists and modelers to identify both the appropriate phenomena and the right level of mod­ eling. Toxicologists can avoid much unnecessary experi­ mentation by using this top-to-bottom approach rather than the currently fashionable bottom-to-top approach.

1.1.4  Analogy to Thermodynamics In physics, Boyle’s law of ideal gases gave rise to thermo­ dynamics, and molecular and mechanistic considerations

led to a theory of gas reactions. The former is based on the idea of finding the minimum number of fundamental vari­ ables that can describe the simplest possible dynamic sys­ tem (P  V  k for ideal gases). The latter required a great deal of knowledge about the mechanism of chemical reac­ tions (wall reaction, activation energy, etc.). Both of these approaches have been attempted in toxicology with, as yet, limited success, as we shall see in subsequent discus­ sions. The reason for the lack of advance in theoretical toxi­ cology is probably that, unlike thermodynamicists, we did not start out by defining the simplest possible toxicological conditions with a minimum number of variables as a point of departure toward more complexity, although coinciden­ tally experiments were conducted under such ideal condi­ tions and in every such instance Haber’s rule proved to be applicable (e.g., Gardner et al., 1977), even though authors may have failed to notice it (Sivan et al., 1984). The lack of conceptualization of the three variables of toxicity resulted in arbitrary study designs, which further eroded the predictability from one experiment to another. It is our opinion that analogous considerations to thermo­ dynamics might help to optimize study design and eventu­ ally to build a theory of toxicology. Thermodynamics like toxicology has three fundamental variables (P, V, and T vs c, t, and W). W (Wirkung in German) will be used for effect, because of the many Es (exposure, elimination, effect, excretion) in English. Before the development of a compre­ hensive theory of thermodynamics, it was clear to scientists that, to study an independent and a dependent variable, a third or other variables had to be kept constant. We have not done this in toxicology, although most dose-response stud­ ies were conducted at constant time (isotemporal). However, to study the relationship between time and effect, the dose needs to be kept constant (isodosic). Moreover, to exam­ ine the relationship between dose and time, the effect must be kept constant (isoeffective). The c  t product will not emerge from the equation of ergodynamics (Wirkungslehre) until after elucidation of the relationship between specific effect at constant time and specific effect at constant dose. In other words, we must learn more about k before signifi­ cant theoretical advance is possible (Fig. 1.2). As mentioned before, most experiments were conducted isotemporally in the past (14 days, 90 days, 104 weeks), which is appropriate for dose-response studies. The arbitrary choice of these time points and the inexactitude of diagno­ sis (stuff them and count them) led to a great deal of confu­ sion in the 14-day studies, because different dose responses, meaning different mechanisms, were often lumped together. Experiments in toxicology have frequently been conducted under isoeffective conditions, mostly with the end point being 100% of an effect (mortality, cancer). However, systematic investigation of c  t  k has not been done, for example, at 20 or 80% of an effect. Finally, there are very few experi­ ments that were conducted under isodosic conditions, because this requires that the concentration be kept constant at the site

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

Ergodynamics (Wirkungslehre) dW = ∂W ∂c

t



Ergodynamics c × t = k × Effect (Wirkung)

dc + ∂W dt ∂t c

Thermodynamics dW = 0 isoeffective – ∂W ∂c

t

P×V=n×R×Τ

dc = ∂W dt ∂t c

dW = ∂W dc + ∂W dt ∂c t ∂t c

dt = 0 isotemporal dW = ∂W dc ∂c t

W.......Effect [Action (Wirkung)] c.........dose t.........time

dc = 0 isodosic

Physics

dW = ∂W dt ∂t c

Action (Wirkung) = Energy × Time

Figure 1.2  Conceptualization of ergodynamics (Wirkungslehre) at constant action (Wirkung), at constant dose, and at constant time.

Toxicology

of action. The only experiment-driven condition, other than aquatic and in vitro toxicology, that keeps the concentration at the site of action constant is continuous inhalation expo­ sure. For example, Gardner et al. (1977) have reported such data after continuous inhalation exposure of experimental animals to benzene and SO2 when the end point in question was measured immediately after termination of exposure (chronaxy, leukopenia). However, when the end point of mea­ surement (streptococcal infection-related mortality) did not occur immediately after cessation of NO2 exposure the time response started flattening out (Gardner et al., 1979). A sys­ tematic investigation of these issues has been done recently for HpCDD after oral administration with as yet only one end point of toxicity (delayed acute toxicity) as end point of measurement (Rozman, 1999), although confirmation of the analysis is emerging for anemia and lung cancer as well (Rozman, 2000b; Rozman et al., 2005). These data pro­ vide support for the suggestion of Rozman et al. (1996) that viewed the dose-time-response as a three-dimensional surface area similar to but conceptually distinctly different from the traditional model of Hartung (1987). Experiments conducted under isoeffective conditions (slices parallel to the dose-time plane) correspond to Haber’s rule of c  t  k represented by hyperbolas. Studies carried out under isotemporal conditions (slices parallel to the time-effect plane) yield S-shaped doseresponse curves along which c  t  k  W whereas isodosic investigations (slices parallel to the dose-effect plane) produce S-shaped time-response curves along which c  t  k  W also. Indeed plotting of the c  t product against W (effect) for HpCDD for doses causing about 10–90% wasting or hemorrhage yielded a straight line (Rozman, 2000a) of high correlation (r2  0.96). This is a beginning core of a theory of toxicology, analogous to P  V  k for isotherms and P  V  k  T for isobars or isochors. Of course ther­ modynamicists know that k  n  R, but ­ toxicology is

Effect (Wirkung) = Dose × Time Figure 1.3  Analogy between ergodynamics and thermodynamics.

not yet there. What is already clear at this junction, how­ ever, is that the dimension of P  V is energy, whereas the dimension of c  t is mass (energy)  time  action, which is called effect in toxicology (Fig. 1.3). The action of a chemical may also be defined as

A  ANS  AS

where AS 

∫c dt



total action (A) consisting of nonspecific action (ANS) plus specific action (AS). The nonspecific action is comparable to heat in thermodynamics that dissipates without being converted to work by an expanding gas. A chemical may have several specific actions (toxicodynamic) such as enzyme induction porphyria and liver cancer. At the same time the organism may have specific or nonspecific actions on the chemical. That portion of a dose which is converted to anything else but the effect of interest must be viewed as nonspecific action with regard to this particular effect. For example, if a chemical is rapidly converted to a much less toxic metabolite and eliminated, then such a com­ pound will have very little specific action. In the case of metabolic activation only the portion of the dose which is converted to the more toxic metabolite will constitute the specific action and the rest must be viewed as nonspecific action. For example a suprathreshold dose of TCDD will be very efficiently converted to specific toxic action as there is very little biotransformation taking place and the toxic moity is TCDD itself. Therefore, in this case kinetics drive the toxicity of TCDD. Dynamics (binding to DNA) is the driving force for the toxicity of nitrosamines and they

Hayes’ Handbook of Pesticide Toxicology

10

require metabolic activation, which is just one of several possible metabolic pathways. Therefore, nitrosamines are less efficiently converted into toxic action than is TCDD. Substituting for c  k/t or for t  k/c and integrating between c1 and c2 or t1 and t2 yields for isoeffective condi­ tions another logarithmic form of Haber’s rule: ln

c2 t  ln 1 c1 t2



Hayes did not make a distinction between chemicals whose action is dominated by toxicodynamic processes and those whose action is determined by toxicokinetic pro­ cesses. The decision tree in Section 1.1.1 (Fig. 1.1) requires identification of the rate-determining (or limiting) step(s). Accordingly, neurotoxicity, teratogenicity, carcinogenicity, hypersensitivity, and induction of enzymes are examples of toxico(pharmaco)dynamic processes, whereas metabolism (absorption, distribution, biotransformation, excretion) and storage are examples of toxico(pharmaco)kinetic processes.

for which an analogy also exists in thermodynamics.

1.2  Kinds of toxicity Toxicity may be classified according to the nature or the duration of the injury involved. Toxicology traditionally has been defined as the science of the study of qualitative and, more important, quantitative aspects of injurious effects of chemicals and physical agents in a subject or in a population of subjects. Paracelsus had already recognized nearly 500 years ago that there is no such thing as nonpoisonous and that the dose alone makes a poi­ son not to be poisonous. Even endogenous body constituents and foodstuffs can be deleterious to an organism if present in excessive quantities over prolonged periods of time. Thus, in addition to the dose, time is the second important variable with which the science of toxicology deals. What then is toxicity? It is the accumulation of injury over short or long periods of time that renders an organism incapable of functioning within the limits of adaptation or other forms of recovery. Therefore, a more appropri­ ate definition of the scope of toxicology would be that it is the science that elucidates the causality chain of interac­ tions and their time course (exposure) between biological entities (subjects) of different intrinsic susceptibility and chemical and physical entities (agents) of different intrinsic potency. Thus, modern toxicology determines in a broader sense exposure responses consisting of dose-responses and time-responses thereby establishing practical thresholds which define the safety of chemicals.

1.2.1  Nature of the Injury The kinds of injury or change that may be produced by chemicals and are known to be of practical importance in certain circumstances are acute and chronic toxicity in the restricted sense, neurotoxicity, teratogenesis, carcino­ genesis, hypersensitivity, metabolism and storage, and induction of enzymes. The dosage-response relationships in these different kinds of toxicity or change are described in Section 1.4. Observed injury may be a direct result of the action of a toxicant or its metabolite(s), or it may be secondary to malnutrition, hormonal alteration, or some other change caused by the compound(s).

1.2.2  Duration of the Injury 1.2.2.1  Factors in the Chronicity of the Injury At least three major independent factors-compound, dosage, and duration of dosing-and a separately measurable depen­ dent factor-storage-are involved in what is often lumped with misleading simplicity under the term “chronic toxicity.” Some compounds are inherently likely to produce chronic effects, which is largely the same as saying that their effects are highly irreversible. In some instances, a single dose not sufficient to produce any immediate effect or perhaps no detectable immediate effect, eventually leads to chronic illness. It is important to realize that there is no necessary rela­ tionship between the number of doses and the chronicity of illness. If a material capable of producing chronic effects is administered repeatedly, the chance that chronic effects will occur is increased, and the chance that only acute poisoning will occur is decreased. However, both acute and chronic effects can occur as part of a single illness. Among the mate­ rials that can produce chronic illness by a single dose are thallium and arsenic (Moeschlin, 1965), triorthocresyl phos­ phate (Smith et al., 1930), or certain carcinogens (Bryan and Shimkin, 1943; Carnaghan, 1967; Magee and Barnes, 1962; Schoental and Magee, 1957). Undoubtedly some other mate­ rials such as lead often would cause both acute and chronic effects if absorbed at a sufficiently large single dose. Other compounds such as potassium cyanide have pro­ duced only acute illness to this date. In other words, the illness caused by cyanide is similar whether it ­follows a sin­ gle large dose or many somewhat smaller doses. If recovery occurs, it progresses at a rate determined by the severity of illness rather than by the number of doses received. The production of persistent effects is not characteristic of cya­ nide, although such effects may follow tissue anoxia of any cause. Some compounds are intermediate to the examples cited in regard to the chronicity of their effects. Chronic poison­ ing cannot be produced by one drink of alcohol, but persis­ tent excessive drinking can lead to chronic organic damage. There is considerable evidence that prolonged excessive intake of sodium in the form of table salt produces chronic hypertension (Meneely, 1966). In these instances, the easy

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

reversibility of the injury finally is overcome by prolonged high dosage. Much confusion would be avoided if the expression “chronic poisoning” were restricted to chronic disease pro­ duced by a chemical or by the chemical changes secondary to radiation or other physical agents. Because chronic dis­ ease may be caused by a single dose and acute poisoning may follow repeated exposure, the duration of exposure ought to be specified separately. Chronic illness (whether secondary to poisoning, infec­ tion, malnutrition, metabolic disorder, circulatory malfunc­ tion, neoplasia, genetic defect, or some unknown cause) is characterized not only by long duration but by certain pathological features, especially scarring and atrophy. It is sometimes implied that chronic illness is nec­ essarily obscure and difficult to diagnose. This simply is not true. Most of the poisoning produced by the alkyl mercury fungicides is both chronic and tragically obvi­ ous. Difficulty in diagnosis is more likely to be associated with very mild, transient illness or with failure to suspect the possibility of poisoning than with any particular set of clinical characteristics. Hayes (1991) recognized the various relationships between manifestation of toxicity and dosing regimen. The decision tree approach in Section 1.1.1 (Fig. 1.1) provides a straightforward explanation for these various constellations. As the biological half-life of a compound increases (relative lack of elimination) or the recovery half-life from a toxic insult gets longer and longer (relative lack of recovery) the distinction between single dose rate and dose (total dosage) becomes blurred. In the limiting case of an infinite biological half-life (for a given life span) or an infinite recovery half-life (relative to life span) a single dose rate will cause the same degree of an effect as repeated dose rates of regular or irregu­ lar frequency. Departures from these limiting conditions will become more and more pronounced with decreasing elimina­ tion/recovery half-lives relative to the time to effect. Hydrogen cyanide is a case in point for an effect in which recovery is the slower and hence rate-­determining process. When the dose is high and the duration of expo­ sure short, not much recovery can occur during time to death and c  t  k will be obeyed strictly. However, with decreasing dose and increasing time to death, recov­ ery will play a greater and greater role in the overall out­ come of toxicity with increasing departure from c  t  k (McNamara, 1976). Alcohol represents the opposite end of the spectrum by causing highly reversible hepatotoxicity. Here it is not recovery from the toxic insult but accumulation of dam­ age that is the slower and hence rate-determining process. Therefore, nearly lethal acute dose rates ingested repeat­ edly but with adequate recovery periods will not cause chronic hepatotoxicity (cirrhosis) because it is the accumu­ lation of damage and not the injury itself that determines the time course of the disease.

11

1.2.2.2  Reversibility The matter of reversibility is subject to several qualifications. Even when a chemical lesion is rapidly and completely reversible, as in the case of thiamin deficiency in its early stages, severe poisoning may lead to irreversible complica­ tions. Also, many compounds have two or more actions, which may differ in reversibility. Finally, the mode of action of many toxicants is unknown. It is, therefore, important to determine whether animals actually poisoned by a particular toxicant are capable of complete recovery or whether they are left with some residual functional or structural injury. As discussed in the preceding section, it is characteristic of some compounds to produce chronic illness (sometimes after a single dose) and of others to produce illness only if dosage is maintained at a sufficiently high level. Section 1.3.2 pres­ ents a quantitative method for recording the tendency of each chemical to produce cumulative effects following repeated dosing, and the related but separable phenomenon of cumu­ lative storage of compounds or their metabolites is discussed later. The effects of many toxicants, including many pesti­ cides, are fully reversible, but because a number of factors may be involved, the possibility of recovery must be tested directly for each compound. Unfortunately, little use is made of the technique of keeping the survivors of the higher doses of ordinary one-dose LD 50 tests for long periods without further dosing in order to observe possible latent effects. This technique is far from new. It has had some use in Great Britain in the systematic testing of pesticides. In fact, the procedure is recommended explicitly in the Pesticides Safety Precautions Scheme Agreed between Government Departments and Industry, issued by the Pesticides Branch of the Great Britain Ministry of Agriculture, Fisheries and Food (1966). The method is simple and capable of revealing the ultimate in irreversibility. It is especially suitable for dis­ covering what Barnes (1968) has called hit-and-run poisons. As already mentioned a single dose of several natural and synthetic compounds have been shown capable of causing cancer and other chronic injury when administered orally or by other routes. The possible reversibility of lesions produced by repeated doses is often neglected also. One can cite examples in which certain morphological changes of the liver have been called cancer without evidence of invasion or metastasis and without any effort to discover whether the changes would regress if dosing was to be discontinued. Such neglect repre­ sents not only poor toxicology but irresponsibility. Reversibility is an important possibility, but is only one among several others whereby an organism can recover from injury. The other possibilities are adaptation and repair as discussed in Section 1.1.1 (Fig. 1.1). Storage of chemicals is related to their kinetics whereas hit-and-run poisons cause-injury by the dynamics of recovery. 2,3,7,8Tetrachlorodibenzo-p-dioxin (TCDD), mirex, hexachlo­ robenzene (HCB), and polychlorinated biphenyls (PCBs)

12

are examples of chemicals that are stored in lipid and/or protein compartments of the body, which is the reason for their very long elimination half-lives. These very long halflives dominate the manifestation of injury caused by these agents. Opposite to these kinetically acting agents are the hitand-run poisons such as warfarin and soman, both of which have very short elimination half-lives but long recovery half-lives; in the case of warfarin, in the form of reversibil­ ity (Nagashima et al., 1969); in the case of soman, in the form of repair (synthesis of new enzyme) (Rozman, 2000a).

1.3  Quantitation of dosageresponse relationships Scientific study of the effects of chemical or physical agents on living organisms requires measurement. A distinction is made between a measurement that involves an agent alone (dose) or one that involves an agent in relation to an organ­ ism (dosage). A 1.0 mg dose of a compound is identical, whether it is administered to a 20-g mouse or a 5000-kg elephant, but the dosages are vastly different: 50 mg/kg for the mouse and 0.002 mg/kg for the elephant. The sus­ ceptibility of different species or even different individuals can be compared precisely only if their body weight is also considered. This does not mean that large animals always require a higher dose than small ones of the same species to mani­ fest the same effect. The tendency in this direction may be obscured by individual variation, particularly because large animals frequently are older or better nourished than small ones and may differ in other ways also. Even so, the sig­ nificance of individual differences of whatever origin can be studied most effectively if dosage rather than dose is considered. The word dosage is properly applied to any rate or ratio involving a dose. Thus the expressions “milligram per kilogram” and “milligram per square centimeter” both des­ ignate a dosage. Dosages often involve the dimension of time [milligram per kilogram per day (mg/kg/day)] but the meaning is not restricted to this relationship. The acute or one-dose ED 50, defined in Section 1.3.1.1 and generally expressed in terms of milligrams of material per kilogram of body weight, is the universally accepted primary way of expressing acute effects of solids and liq­ uids that are swallowed, contaminate the skin, or are admin­ istered subcutaneously, intravenously, or by other parenteral routes. An LD 50 is a special case of an ED 50 in which the effect measured is death. The numerical form of these ED 50 or LD 50 values permits useful comparisons between the acute effects of different compounds or of the same com­ pound administered by different routes. The 90-dose ED 50 (or LD 50) and the chronicity index both may help

Hayes’ Handbook of Pesticide Toxicology

to express the results of repeated dosing. Finally the same mathematical procedure can be applied to studies of any duration, including lifetime studies. The use of ED 50 and LD 50 values-whether for 1 dose or 90 doses-is the ideal way to express toxicity because these values are direct measures of the dosage received by fish or other organisms that live in water and obtain their oxygen from it. Under these circumstances, the investiga­ tor often must be satisfied with a statistical estimate of the time required for a given concentration to produce a speci­ fied effect in a fixed time. No matter what the physical form of the chemical or the habits of the test species are, there is obvious inter­ est in determining the largest dosage or concentration that produces no observable effect or no significant observable effect. This is the largest safe dosage for the test organ­ ism under the conditions of the test. Such a no-effect-level (NOEL) is often used as a basis for estimating a lower value considered safe under more varied conditions, including the exposure of other species, especially humans (Sections 1.2.7.4 and 1.5.9.1). Finally, dosages may be compared in terms of tissue levels no matter what the physical form of the compound, the habits of the species, the route of absorption, or the duration of dosing. The dosage-response relationship is the most funda­ mental single principle in toxicology. It extends to all kinds of injurious effects (Section 1.4) and implies the existence of a threshold dosage for each compound below which, under defined conditions, no harmful effect is produced (Section 1.3.7.4). Hayes defined dose as mass (mg) and dosage as concen­ tration (mg/kg). This is a useful distinction to generalize the “dosage-response” across species (e.g., mammals vs fish) after different routes of administration (e.g., oral vs ­dermal). It is also advantageous when applied to the concept of Hayes’ index, which is an ingenious attempt to incorporate time as a variable of toxicity without designating time as an explicit function of toxicity. However, it is less accommo­ dating for the herein-developed theory of toxicology, which uses both dose and time as explicit variables of toxicity. The most profound difference between the two approaches arises for safety and risk assessment. If toxicity is viewed as being solely a function of dosage as done by Hayes and others, then the logical consequence is to look for a noobserved-effect level (NOEL) or a lowest-observed-effect level (LOEL) as a point of departure for determining a safe dose. However, considering both dose and time as explicit functions of toxicity leads to having to incorporate both of these variables into safety and risk assessment as suggested by Rozman (2000a). The c  t concept provides a scien­ tifically valid and firm departure point for safety and risk assessment instead of the inherent fuzziness of a NOEL or LOEL.

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

1.3.1  ED 50 or LD 50 1.3.1.1  One-Dose ED 50 or LD 50 An ED 50 is a statistical estimate of the dosage of a mate­ rial that would produce a specified effect in 50% of a very large population of a test species under stated conditions, for example, a single oral dose of an aqueous solution given to male rats. Of course it would be impractical to use hundreds or even thousands of animals to make such a test. Even if this were done, it would be unlikely that the investigator would find the dosage to produce the effect in exactly half of the animals. That is why the parameter must be estimated statistically. In practice, test animals are divided into groups of moderate size, frequently about 10 per group. Each group is given one of a series of geo­ metrically increasing dosages selected in such a way that the smallest dosage will produce the intended effect in only a small proportion of the group receiving it, whereas the largest dosage produces the same effect in the majority of animals receiving it. The result for each group is expressed as the percentage of animals showing the effect under study. By one technique or another, the percentage effect for each group is converted to a probit and related to the logarithm of the dosage that produced it. Any effect measured in this way must be recorded as an all-or-none response. However, phenomena that show continuous variation may be treated on an all-or-none basis merely by selecting an arbitrary limit. For example, systolic blood pressure may vary widely but could be made the basis of an ED 50 by counting all animals whose pressure exceeded 150 mm Hg. Acute toxicity studies for purposes of approximating an LD 50 are conducted much less frequently today than in years past. However, acute lethality studies, if conducted properly, can provide a considerable amount of basic information about the toxicity of a chemical. Clearly, the amount of information obtained will depend on the quality of study-what Boyd (1972) called “cage-side observation.” From the standpoint of accidental exposure to pesticides, one of the most useful pieces of information is the compar­ ison between the dermal and oral acute lethalities. Because occupational exposure to pesticides is largely by the der­ mal route, those pesticides that are absorbed through the skin in sufficient amounts to kill animals that receive small dosages definitely represent more of a potential hazard to people than materials that are not absorbed through the skin in toxic amounts except at high dosages. The one-dose ED 50 or LD 50 has served a useful pur­ pose in defining the approximate toxicity of chemicals with almost no theoretical justification for the way such experiments were and still are conducted. The usual pro­ tocol entails the administration of a compound by some route of exposure and then to determine the effect (death) after a specified length of time, which usually is 14 days. There are numerous and severe problems with this protocol

13

from the theoretical point of view. The 14-day observation period is entirely arbitrary and lacks scientific rationale. It is the time response at constant dose that determines the length of the observation period needed after a single dose Rozman (1999). It can be 70 days (e.g. HCDD-induced wasting) or just a few minutes (e.g. CO-induced asphyxi­ ation) depending on the time to effect (death). Similarly problematic is the counting of dead animals. As strange as it might sound, one dead animal may not be the same as another dead animal if the two died by a different mecha­ nism of action which is often the case after supralethal doses when recovering animals still succumb to a dam­ age not repairable during the allotted observation period (e.g. 14 days). For example, rats exposed to NO2 acutely by inhalation die of either spasm of the larynx or edema of the lungs. Spasm of the larynx and lung edema have very obviously different mechanisms of action and as such are part of different dose and time responses. Therefore, they will distort the c  t concept if lumped together. The biggest problem with the ED 50 and LD 50 studies is that for the most part they were and are not conducted under conditions of toxicodynamic and/or toxicokinetic steady state and, therefore, almost all studies measure not only toxicity, but toxicity and recovery at the same time. Depending on the ratio between recovery/elimination half-life and the observa­ tion period, the dose response will be increasingly distorted in the form of a flattening of the sigmoid curve. Under very unfavorable experimental conditions when very little toxicity versus a great deal of recovery is being measured the lower part of the S-shaped curve appears nearly linear with dire consequences for the accuracy of any risk assessment based on such faulty interpretation. Shape of the ED 50 or LD 50 Curve  Several matters regarding determination of ED 50 and LD 50 values are illustrated by Fig. 1.4 based on LD 50 studies of DDT. All parts of the graph represent the results of tests in which the groups of rats were given various dosages of the compound. In part A, the dosage for each group has been plotted on plain graph paper against the percentage of mortality. The fundamental statistical principles illustrated by part A first were defined clearly by Trevan (1927). Specifically, he pointed out that there was no such thing as a minimal lethal dosage or minimal effective dosage conceived at the time, namely, a dosage that would be just sufficient to produce the effect in all animals of a given species. He noted that the variability of individuals in a population led to the char­ acteristic S-shaped curve and that there seemed to be less variability at the 50% level of response. Trevan proposed the equivalent terms “median lethal dose” and “LD 50,” both in their presently accepted meaning. He also sug­ gested that dosages that kill other proportions of large groups of animals be designated by analogous symbols, for example “LD 25” and “LD 75” for dosages that kill

Hayes’ Handbook of Pesticide Toxicology

14

Mortality (%)

100

A

B

75 50 25 0

0

40

80

120 160 200

20

50

100 200

Dosage (mg/kg) 99.9 99

8 C

D

E 7 6

50

5

Probits

Mortality (%)

90

4

10

3 1 0.1 30 50

100

30 50 30 50 100 Dosage (mg/kg)

100

2

Figure 1.4  Mortality of white rats caused by oral doses of DDT. Each point represents one group of animals, usually 10; (A) percentage mortality in six separate tests plotted against dosage; (B) the same data with percentage plotted against logarithm of dosage; (C) the same data with percentage mortality expressed as probits plotted against logarithm of dosage; the 19/20 confidence limits are shown by dotted lines on either side of the dos­ age–response curve; (D) dosage–response curves for each of the six separate tests; (E) the dosage–response curve that differed most from the others, showing its relatively wide confidence limits.

25 and 75% of the group, respectively. In Trevan’s paper in the Proceedings of the Royal Society, the symbols were printed with a space between the letter “D” and the appropriate number. This style has been adopted for this chapter, partly because it is authentic and partly because it is completely clear when it becomes necessary in theoretical discussions to refer to fractions of a percentage, for exam­ ple, “LD 0.01.” The LD 50 can be read from the curve even in its S-shaped form on plain graph paper. Thus, in part A, the level for 50% mortality intersects the curve at a dosage of about 113 mg/kg. It may be noted parenthetically that the middle portion of the sigmoid curve-in the region of

20–80% response is often indistinguishable from a straight line. The fact that a simple straight line relationship between dosage and percentage response adequately describes some sets of data must not obscure the fact that more complete data determine a sigmoid curve on plain paper. Part B of Fig. 1.4 represents the same data shown in part A, but dosage is now shown on a logarithmic scale rather than on a simple arithmetic scale. The S-shaped pat­ tern persists, but the curve approaches a straight line. Part C of the graph represents the same data plotted with an addi­ tional conversion. Here the logarithm of dosage is plot­ ted against percentage mortality expressed as probits. The correspondence between percentage and probits is shown

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

by the scales on the left and right of the lower portion of the graph. As may be seen, the points are scattered about a straight line when the full (logprobit) conversion is made. The logarithmic conversion apparently was introduced first by Krogh and Hemmingsen (1928), but it was the subject of numerous publications cited by Gaddum (1933) in the classical paper in which he introduced the full conversion essentially in the form still used today. Actually, Gaddum employed normal equivalent deviations rather than the probits that are commonly used today. However, a ­ probit is merely a normal equivalent deviation to which 5.0 has been added for convenience to eliminate negative values. The paper by Gaddum was shortly followed by three papers by Bliss (1934a,b, 1935a). These four papers on the sta­ tistical relationship between dosage and response are still basic today. The facts regarding probit analysis were sum­ marized in a masterful way by Finney (1971) in a book first published in 1947 and revised in 1971. A more gen­ eral treatment of the principles of biological assay is that of Emmens (1948). It must be stated that the slender sigmoid curve we have discussed constitutes (insofar as data permit one to judge) a cumulative lognormal curve. Conversion of the percentage response to normal equivalent deviations or probits is merely a statistical device for converting the sig­ moid curve into a straight line. Although Gaddum (1933) employed the logarithmic conversion as well as the nor­ mal equivalent deviation conversion for use in toxicity measurements, it apparently was not until 1945 that he introduced the word “lognormal” to describe the situation in which log x is normally distributed. In the same paper, Gaddum (1945) emphasized that the distribution of values for many parameters in nature is not statistically normal. This means only that the distribution frequently does not conform to any of the family of curves commonly called normal or Gaussian after Karl Friedrich Gauss, who first popularized this particular pattern of variation. Use of the word “normal” in this connection has no bearing on physi­ cal or biological normalcy. In fact, Gaddum pointed out that if the distribution of the volume of particles is normal the distribution of their diameters will, of necessity, not be normal. Gaddum stressed the importance of converting mea­ surements in such a way that the results may be subjected to statistical evaluation. In addition to logarithmic conver­ sion of each variable for this purpose, he suggested that a positive or negative constant might be added to each vari­ able prior to its logarithmic conversion. The logprobit conversion has great value for purposes of description and statistical analysis. However, in spite of its great practical value, the basic assumption that the rela­ tionship of variables is perfectly lognormal cannot be con­ sidered proved, because the upper and lower extremities of the curve have not been studied experimentally to a suffi­ cient degree. This detail is discussed in Section 1.3.7.4.

15

The shape of the dose-response curve has been the subject of vigorous discussions because of its utmost importance for risk assessment. Statistical approaches for its description abound but a theoretical treatment of it had not been proposed until Rozman et al. (1996) suggested, as a first simplified approach, a Malthusian-type statement of a problem described by the Verhulst equation address­ ing the issue of change under constraint. The constraint in toxicology is the impossibility of having more than 100% of a population affected. Thus, the effect as a function of dose or time is always proportional to the effect that has already occurred as well as to the effect still remaining to occur. The Verhulst equation has an exponential solu­ tion in terms of effect and not in terms of dose, which is compatible with the notion that the effect in a population has logarithmic properties (normal distribution) but the dose (number of molecules) does not have such properties. However, the dose-response has been traditionally plot­ ted on a log (dose) vs effect (arithmetic) scale. Because log (dose) vs effect is the inverse function of log (effect) vs dose (arithmetic) and because inverse functions are entirely symmetrical, there has been no problem with this plot, even though in terms of epistemology the traditional way of plotting the dose-response function may not nec­ essarily be entirely correct. It must be emphasized, that nonlinearity of the dose-response can also be derived from thermodynamic considerations as well (Rozman, 2003 a,b; Waddell, 2008). Under ideal conditions of toxicodynamic or toxicokinetictoxicodynamic steady state any dose response is extremely steep, best exemplified by inhalation anesthesia (see Storm and Rozman, 1998). There is a factor of no more than 2 between doses, which will anesthetize the most and least sensitive individual. Inhalation of a volatile agent is kineti­ cally related to intravenous infusion, which for compounds of short half-life will provide steady state concentrations very rapidly. Similarly, the dose causing 100% wastinghemorrhage and 0% of this effect in rats under conditions of toxicokinetic steady state is also a factor of 2 (Rozman, 1999). Similar considerations are valid for compounds that act by toxicodynamic mechanisms. Departure from either type of steady state condition will introduce recovery as an additional variable. The vast majority of toxicological exper­ iments (and real-life situations) are not conducted or do not occur under steady state conditions. The resulting introduc­ tion of one or more variables in addition to toxicity is the reason for the large variability in interlaboratory experi­ ments and the mistaken assumption of flat dose-responses. Under ideal conditions, when all variables other than toxic­ ity are controlled, all dose-responses are as steep as the ones discussed previously. ED 50, ED 1, ED 99, and Corresponding LD Values  Returning to part C of Fig 1.4, it may be seen that the level of 50% mortality intersects the curve at a dose of 113 mg/kg.

Hayes’ Handbook of Pesticide Toxicology

16

This is the oral LD 50 for DDT as indicated by the observed data. In a similar way, the 1% mortality level intersects the curve at a dosage of 52 mg/kg, which, therefore, is the LD 1. Mortality of 99% is not shown on the graph but would fall at a dosage of 223 mg/kg, which is the LD 99.

95

Confidence Limits and Reproducibility  The degree of scatter of observed values may be evaluated by calcula­ tion and expressed as a confidence limit. These limits are shown by dotted lines on both sides of the solid line in part C of the graph. These particular confidence limits indicate the area or range within which the dosage-response line may be expected to fall in 19 of 20 samples taken at ran­ dom from the same population. It may be seen that a series of such curves will correspond closely with one another at the 50% mortality level but will agree less well as the mor­ tality approaches either 0 or 100%. This is a graphic rep­ resentation of the fact first noted by Trevan (1927) that the LD 50 may be estimated more accurately than correspond­ ing statistics for greater or lesser effect (e.g., LD 99 or LD 1). This is also true of ED 50 values and corresponding ED 99 and ED 1 values. The points represented in parts A, B and C of Fig. 1.4 represent the results of six separate tests made in six dif­ ferent years for the purpose of determining whether there was any change, genetic or otherwise, in the susceptibil­ ity of the particular colony of rats to acute poisoning by DDT. Part D of the graph shows the dosage-response lines determined in connection with the six separate tests. The lines correspond very closely at the 50% mortality level but diverge somewhat in connection with higher or lower mortality rates. Actually, all of the lines are in good agree­ ment, indicating that there was no detectable change in the colony concerning susceptibility to DDT. In fact, the dependability of this kind of test is well recognized. Weil et al. (1966) reported that they had done one-dose oral LD 50 of 26 chemicals annually for 11 or 12 years to deter­ mine the reproducibility of the test and the dependability of commercial production of the chemicals. The resultant median lethal doses were relatively unaffected by the dif­ ferent annual samples of each chemical, by changes in the stock or rats, by the degree of dilution of the toxicants, or by change in the personnel performing the tests. Only one variable, the weight of the rat, appeared to have a sig­ nificant effect on the values obtained, which is consistent with a report of Lamanna and Hart (1968) as interpreted in Section 1.5.1 by Rozman and Doull. Slope and Its Relation to Confidence Limits  Part E of Fig. 1.4 shows the data and resulting curves for a single LD 50 determination, namely, the particular test that dif­ fered most from the average of the six tests. It may be seen that the slope of the line is greater than the slopes of the other LD 50 lines (part D). This increase in slope

Mortality (%)

90 70 50 30 20 10 5 1 0.1 10

20

30 50 100 Dosage (mg/kg)

200 300

Figure 1.5  Dosage–response curves for dieldrin (O) and toxaphene ( ) given orally to white rats. Points adjusted according to the method of Litchfield and Wilcoxon (1949) are distinguished by a superimposed x.

•

is a graphic representation of the data on which this par­ ticular determination was based. The greater ­variability of the data is also reflected by the fact that the dotted curves (representing confidence limits) lie farther from the solid line than the corresponding curves do in part C of the figure. Procedures for Determining ED 50 and LD 50 Values  Parts C and E of Fig. 1.4 represent actual determina­ tions of LD 50 values using the graphic method proposed by Litchfield and Wilcoxon (1949), who also supplied details of the method for calculating the 19/20 confi­ dence limits. A number of non-graphic methods are avail­ able for determining ED 50 or LD 50 values, including the methods of Bliss (1935a,b, 1938). The non-graphic methods have in common the fact that percentage values must be ­ transformed by means of an appropriate table or ­calculation. A wide variety of methods have been reviewed by McIntosh (1961), who concluded that the differences among results with the 15–85% response range are negli­ gible. Thus, selection of the method to use depends largely on personal choice. Repeated determinations of an ED 50 or LD 50 for a particular compound under the same conditions should give not only statistically indistinguishable values but also statistically indistinguishable slopes of the dosage-response curves as shown in Fig. 1.5. The curves may be related in such a way that the ED 50 (or LD 50) values are statis­ tically distinguishable but other values such as the ED 1

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

(or LD 1) are not distinguishable. Figure 1.5 offers an example. The LD 50 values for dieldrin and toxaphene are different but the LD 1 values for these compounds are sta­ tistically indistinguishable. Hayes was well aware of the problems arising from a lack of control of variables. He, like others, proposed a sta­ tistical treatment of data to deal with “hidden” variables, which cannot be readily identified. This was appropriate at the time he wrote this chapter. However, even a partial development of a theory of toxicology requires a different type of approach. The question is not how to accommo­ date the apparent difference in the slope of the dieldrin and toxaphene mortality dose responses statistically. Rather the question is why is the toxaphene dose response slope flat­ ter than that of dieldrin (Fig. 1.5). The half-life of dieldrin in rats is estimated in the range of weeks (Mueller et al., 1975) whereas that of toxaphene is in the range of 4–7 days (ATSDR, 1996). Therefore, during an observation period of 14 days little of dieldrin will be eliminated whereas most of toxaphene will have been excreted. Consequently, not much recovery takes place with dieldrin during the obser­ vation period (14 days), whereas a great deal of recovery occurs with toxaphene combined with the concurrent devel­ opment of toxicity. It must be understood that just because some animals eventually succumb does not mean that they did not try to recover from the damage if time for recovery was available relative to the half-life of the compound or of the effect. If toxaphene had been administered in a way to make its kinetics similar to that of dieldrin (loading dose rate followed by maintenance dose rates) then the two dose responses would have been parallel (identical slopes). This is an extremely important issue. It is clear that dieldrin is a more toxic compound than toxaphene, which is also indi­ cated by Fig. 1.5 when comparing the highest doses of these two chemicals. However, comparing the lowest doses suggests that at low doses toxaphene is equally as toxic as or more toxic than dieldrin. This is of course impos­ sible because relative potency is an intrinsic property of a chemical. In fact, this illusion of dose-dependent relative potency arises as a result of a lack of experimental con­ trol of the kinetics of toxaphene, whereas that of dieldrin is controlled coincidentally by its long half-life. The lack of kinetic control is very frequently one of the “hidden” variables in toxicologic experiments. Similarly, a lack of control of absorption after dermal application of chemicals can be only partially controlled by standardizing dosing volumes. Applied dermal dose will still remain a poor sur­ rogate for systemic dose in most instances as discussed in Section 1.3.6. Test Using Small Numbers of Animals  The conventional procedures for determining ED 50 or LD 50 values require the use of approximately 50–100 animals. The use of this many rats or mice may be practical, even though somewhat expensive. The use of a similar number of dogs or monkeys

17

often is entirely impractical. To meet this problem a num­ ber of methods have been developed that permit the use of a small number of animals per group to determine approxi­ mate ED 50 or LD 50 values (Gaddum, 1933; Deichmann and LeBlanc, 1943; Weil, 1952; Smyth et al., 1962). Volume of Each Dose  If the results of toxicological tests are to be compared, it is wise to keep all conditions as nearly uniform as possible. One variable that should be stan­ dardized is the volume of solution or suspension in which compounds are administered. It has been found practical to give most oral doses at the rate of 0.005 ml per gram of body weight and to give dermal applications at the rate of 0.0016 ml per gram of body weight. Differences in dosage are determined by changing the concentration. The value of 0.0016 ml/g was chosen for dermal application because it represents a plausible exposure of about 100 ml for a human and also gives even numbers for dosage associated with many formulations actually used in the field. Thus, at this rate of application, the dose for a 70-kg human would be 112 ml – a not unlikely degree of contamination as the result of spillage. Formulations of 0.312, 0.625, 1.25, 2.5, 5, and 10% produce dosages of 5, 13, 20, 40, 80, and 160 mg/kg, respectively, when applied at the rate of 0.0016 ml per gram of body weight.

1.3.1.2  90-Dose ED 50 or LD 50 It has been suggested, more or less empirically, that sub­ acute tests should occupy up to one-tenth of the life span of the experimental animals (commonly considered to be about 90 days for the rat and 1 year for the dog) [Food and Agriculture Organization/World Health Organization (FAO/WHO), 1958]. Boyd (1961) accepted the concept of one-tenth of the life span, but considered it to be 100 days in the rat. However, the important thing is not the choice or definition of a particular fraction of the life span but the selection of a testing interval that is as short as practicable and yet will give meaningful information about the effect of absorbing the toxicant during an entire lifetime. Secondarily, it would be desirable to have a standard test so that results from dif­ ferent laboratories would be reported in the same terms. Apparently Boyd and Boyd (1962) were the first to report subacute toxicity in the form of an LD 50. The com­ pound was administered intramuscularly for as long as 100 days. In connection with oral doses it was proposed (Boyd and Selby, 1962) that the compound under test be administered by stomach tube for 100 days. The test dif­ fered in some technical requirements from the 90-dose test described subsequently. In spite of this, the two tests are fundamentally similar, and the results of one are largely interchangeable with those of the other. Several years after he had proposed the 100-day test, Boyd (1968) pointed out that, for compounds he studied,

Hayes’ Handbook of Pesticide Toxicology

18

99.99 99.9 99

Mortality (%)

95 90 70 50 30 10 5 1 0.1 0.01 0.01

0.03 0.05 0.1

0.3 0.5

1

2 3

5

10

Dosage (mg/kg/day)

•

Figure 1.6  Oral dosage–response curves for 1-dose (O) and 90-dose ( ) tests of warfarin. Points adjusted according to the method of Litchfield and Wilcoxon (1949) are distinguished by a superimposed x. From Hayes (1967b), by permission of Academic Press.

Table 1.1  Toxicity of Warfarin to Male Ratsa 1-Dose

90-Dose

Number of rats tested

50

110

Survival time (days)

5–10

3–43b

1.6

0.077

1.4–1.9

0.055–0.108

1.0

0.04d

0.84

0.032

LD 50 19/20 confidence limits Lowest dose to kill LD 1c

c

c

a

From Hayes (1967b), by permission of Academic Press. The true range may be 3–25 days; warfarin was probably not the cause of death in the rat that died after eating the compound for 43 days. c Expressed as milligrams per kilogram for 1-dose test or milligrams per kilogram per day for 90-dose test. d This death probably was not caused by warfarin. The smallest dosage to cause a death clearly related to warfarin was 0.08 mg/kg/day. b

the test could be reduced to about 70 days with little or no loss of important information. Weil and McCollister (1963) showed that the results of 90-day studies not only in rats but even in dogs were simi­ lar to corresponding lifetime studies in these species for a wide range of compounds. Hayes (1967b) pointed out that a 30-day test in the rat would be adequate for some compounds (e.g., potassium cyanide). However, a review of data on certain chemosterilants showed that, although 30 doses were entirely inadequate to reveal the potential injury caused by repeated doses of any

of them, 90 doses gave for most of them essentially the same results as those of tests lasting twice as long. Thus, although a test involving fewer than 90 doses would be adequate to reveal the effect of long-term exposure to many compounds, a 90-dose test is more generally valid for predicting life­ time effects. In fact, even a 90-dose schedule is inadequate to define the long-term toxicity of some compounds such as hempa (hexamethyl phosphoric triamide); however, it was clearly evident after 90 or even somewhat fewer days that animals being dosed with hempa were still dying and that the exposure would have to be prolonged to assess the toxicity properly. Thus a standard of 90 doses was selected for quantitative study of the effect of repeated doses, partly because 90-dose tests were already widely accepted for other purposes and in spite of the limitations just mentioned. The 90-dose ED 50 (or 90-dose LD 50) is statistically comparable to a 1-dose ED 50 (or 1-dose LD 50). In both tests, percentage effect expressed as probits is related to dosage expressed as logarithms. The results of a 1-dose and a 90-dose LD 50 study of warfarin are shown in Fig. 1.6. It may be seen that the curves are similar in slope although the dosages that proved critical differ by a factor of about 20. Table 1.1 shows the LD 50 values and related statistics for warfarin. Hayes (1991) as well as others struggled between practicality and awareness of the importance of time in multiple-dose-rate experiments (subacute, subchronic). Unfortunately, a theory leaves no room for considerations of practicality because laws of nature exist on their own timescales with complete disregard for human convenience. As was the case for acute studies with the 14-day obser­ vation period the timescale of the 90-dose-rate studies is largely arbitrary. Naturally, a number of effects will become manifest in 90 days (plus an additional 14-day observa­ tion period for recovery) that cannot be seen in 14 days. Nevertheless, the selection of 90 days or 104 weeks is as arbitrary as the 14-day observation period after single dose rates or after the 14-day off-dose observation period at the end of a subchronic experiment. A similarly arbitrary tim­ escale is the l04-week carcinogenicity bioassay. Having recognized time as a quantitative and quantifiable variable of toxicity together with the need for mechanistic defini­ tion of an effect requires that each effect must be studied on its own timescale. Although the 90-dose-rate study may be still retained for the time being as a rough first estimate of potential sub-chronic effects, mechanistic studies should be conducted on the timescale of a given effect. For exam­ ple, sub-chronic warfarin-induced hemorrhagic death does not need to be studied in 90-dose-rate studies because no treatment-related lethality occurred after day 25 (Hayes, 1967b). Clearly recovery in the form of adaptation has taken place. Any further sub-chronic studies should have been conducted on a timescale no longer than 25 or per­ haps 30 days, if the population studied was no larger than that used by Hayes (1967b).

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

99.99 99.9 99 95 Neoplasms (%)

Determination of the 90-Dose ED 50 or LD 50  In calculating the conventional 1-dose LD 50, no account is taken of time, although, of course, animals given a single dose do not respond to it simultaneously. The 90-dose ED 50 is managed in a similar way; the animals are held long enough after the last dose to be sure that all reactions have been counted. The two procedures differ, because in the 1-dose test all animals in a group receive the same dos­ age, but in the 90-dose test animals in the same group may receive different total dosages because some may survive longer than others. This difference does not invalidate a comparison of the results of the two tests, but makes quan­ tification difficult (Saghir et. al., 2005). In determining the acute oral ED 50, the compound is usually administered by stomach tube. In determining the oral 90-dose ED 50 the compound is administered as a mix­ ture in the diet. This difference in technique for oral admin­ istration introduces a second kind of difference (applying to this route only) between the two kinds of oral ED 50 val­ ues, but it has two advantages: convenience and realism. Obviously it is more convenient to maintain an animal on a special diet for 90 days than to dose the animal by stom­ ach tube for the same period. Except in connection with drugs, which are administered in discrete doses, it is also more realistic to administer repeated doses in the diet rather than by stomach tube. If people receive relatively regular repeated doses of an environmental compound, the intake is usually distributed throughout a considerable portion of most days while the result of a single massive exposure due to a splash or other spillage associated with occupation, or the ingestion of a relatively concentrated material due to accident or suicide. Furthermore, if a number of compounds are administered by any particular route in such a way that the absorption of a single dose of each is concentrated in as short a period as possible, but the absorption of repeated doses of each is distributed as evenly as is practical over each day, then any difference in cumulative effect of the compounds will be demonstrated to greatest advantage. Thus, to determine the oral 90-dose ED 50 (or LD 50) of a compound, appropriate concentrations in ground chow are fed to groups of animals for 90 days. All survivors are then fed chow without the compound for a minimum of an addi­ tional 2 weeks, and, if any of them are still affected, they must be observed for as long as necessary until they have died or recovered. The dosage (expressed as ­milligrams per kilogram per day) is calculated from measured food con­ sumption (Hayes, 1967b). Although Hayes (1991) and others were aware of the importance of time, time was dealt with only semiquanti­ tatively in toxicology. It is true that in the single-­dose-rate studies the animals receive the same dose whereas in ­multiple-dose-rate experiments they may receive different doses because animals die at different times. It is also cor­ rect that this does not invalidate the comparison between them because the concentration at the site of action (steady

19

80 50 20 5 1 0.1 0.01 10

30

60 100 300 2-AAF (ppm)

600 1000

Figure 1.7  Bladder neoplasms in dead, moribund, and sacrificed mice fed 2-AAF continuously: ( ) 33 months; () 24 months; (O) 18 months. The graph is based on data by Farmer et al. (1980).

•

state) is the critical variable, for which the dose or dose rate is just a surrogate measure. Administering a compound by gavage in the 1-dose-rate studies or mixed with the diet in the 90-dose-rate stud­ ies again has a practical basis. The absorption phase after gavage provides for a longer near steady state condition for compounds having half-lives of about a day or longer. Feeding a compound to rats also improves kinetics in the course of subchronic experiments, because rats have two or more feeding periods per day, which provides more nearly steady state conditions for compounds of intermediate halflife than does a single dose rate per day by gavage. It is unfortunate that the widely used study designs (acute, sub­ chronic, chronic) in toxicology were developed with little or no consideration of kinetics. This is understandable, though, in historical context because study designs of toxicological experiments were already firmly ingrained by the time the first book was published on pharmacokinetics (Dost, 1953). Nevertheless, no significant advance in theoretical toxicology can take place until study designs are changed to accommodate toxicodynamic and/or toxicokinetic time scales as quantitative and quantifiable variables of toxicity.

1.3.1.3  The ED 01 and Related Studies Logprobit analysis may be applied to any study of dosage response regardless of the duration of dosing or the effect that is recorded. The following paragraphs outline this kind of analysis of two studies of cancer. The theoretical basis for using logprobit analysis to investigate the fundamental problem of small dosages is discussed in Section 1.3.7.4.

Hayes’ Handbook of Pesticide Toxicology

20

99.99 99.9 99

Neoplasms (%)

95 80 50 20 5 1 0.1 0.01

10

30

60 100 300 2-AAF (ppm)

600 1000

Figure 1.8  Liver neoplasms in dead, moribund, and sacrificed mice fed 2-AAF continuously: ( ) 33 months; () 24 months; (O) 18 months. The graph is based on data presented by Farmer et al. (1980).

•

The ED 01 Study  As of 2009 there had been only one sta­ tistical study for which the raw data were easily available of the effects of small dosages of a chemical. In this ED 01 study (Staffa and Mehlman, 1980; Hughes et al., 1983), a predetermined number of mice at each dietary level were killed, after being fed continuously at different dietary lev­ els of 2-acetyl-aminofluorene (2-AAF), at intervals of 9, 12, 14, 15, 16, 17, 18, 24, and 33 months. One paper in the original report (Farmer et al., 1980) presented the raw data for tumor incidence and also pre­ sented logprobit graphs of the incidence of bladder tumors as related to dietary concentration for the 18-, 24-, and 33-month intervals. A similar graph was presented for the incidence of liver tumors. These two graphs were certainly unusual in that the points on which they were based were not shown. Corresponding graphs have been prepared that differ in two ways: the observed points have been plotted and the control values have been noted. Values that could not be plotted on probit paper-because the expected values are too small or too large to be subject to correction accord­ ing to the method of Litchfield and Wilcoxon (1949)-have been shown by arrows indicating the direction at which the points would lie at infinite distance to the left on logarith­ mic paper. The graph for the bladder (Fig. 1.7) confirms what the original authors admitted, that is, that the curves for different time intervals are consistent with the view that there is a threshold dosage below which 2-AAF does not increase the incidence of bladder tumors above that seen in controls. In other words, the curves remain straight until they reach the area of the graph where the incidence values

are indistinguishable from those observed in the controls. It could be argued that some of the values at the lower part of the curve tend to be displaced upward and to the left. However, this interpretation is unjustified because the inci­ dence of tumors in the control group showed no trend in terms of time. As a matter of fact, the highest incidence was observed at the first interval, that is, a value of 1.47% at 9 months. Thus, the entire range of incidence observed among the control groups must be considered. If this is done, the points that might otherwise be interpreted as occurring above and to the left of the straight line are seen simply to be indistinguishable from control values. Farmer and his colleagues (1980) argued that the graph for liver tumors completely excluded the possibility of any threshold. In the summary and conclusions (Gaylor, 1980) it was stated that “liver tumors showed a nearly linear response over the experimental dose range, thereby dispel­ ling any notion of a threshold dose.” One would have to agree with this statement if it was intended to mean that the results ruled out a threshold at any one of the dietary ­levels of 30 ppm or higher. However, Fig. 1.8 shows that the curve did closely approach the control levels and suggests that, if a wider range of dosages had been used, the thresh­ old might have been encountered in the range of 16–23 ppm or, more conservatively, in the range of 10–30 ppm. The first report on the ED 01 study is notable for neglecting the control values and, more broadly, for ignor­ ing the scientific question of whether the values observed above the noise levels determined by the controls were consistent with a cumulative lognormal distribution. The report of the workshop held in September 1981 (Hart et al., 1983) was apparently concerned mainly with integrating time with dosage-response. It was concluded (Hughes et al., 1983) that “the ED 01 Study demon­ strates the observed risk is more adequately expressed in a time and dose continuum rather than simply as dose.” The conclusion undoubtedly is correct and was supported by very sophisticated calculations. However, the question of whether the observed increase in incidence above con­ trol values was cumulative lognormal in distribution was neglected. The conclusion was that “even with a study as large as the ED 01 study, statistical uncertainty makes it impossible to establish the true shape of the dose-response curve at low tumor rates. Neither can such studies prove or disprove the existence of thresholds.” The ED 01 study, also called the megamouse experi­ ment, is indeed very important for toxicology. It provides several lessons beyond those identified by Hayes (1991). The ED 01 study was designed with a lack of kinetic con­ siderations but the initiators were lucky that the dynamic half-life of 2-AAF-induced damage is such that feeding it in the diet provided a steady state of injury (Rozman et al., 1996; Rozman, 2000b). It was also fortuitous that there was not enough toxicity as the experiment progressed

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

establish a complete dose response for that particular end point of toxicity. Nevertheless, it is possible to construct a hypothetical dose-response curve beyond the natural life span of a species using the c  t  k  W conversion. Such a hypothetical dose response for 2-AAF becomes also very steep for liver cancer beyond the natural life span of mice. It is important to note that this biological interpre­ tation of the ED 01 study is entirely consistent with sta­ tistical approaches as exemplified by the Hartley-Sielken model using both dose and time as variables of toxicity (Hartley and Sielken, 1977a,b), but is completely at odds with currently used linear extrapolation models using only dose as a variable of toxicity.

99.99 99.9 99 Neoplasms (%)

95 80 50 20 5 1 0.1 0.01

21

1

10

100

1000

Dosage (rads) Figure 1.9  Incidence of thymic lymphoma (O), myeloid leukemia (), reticular cell sarcoma (), and totalreticular cell neoplasms ( ) in mice that received different dosages of gamma radiation. The graph is based on data presented by Ullrich and Storer (1979a).

•

and therefore the 24- and 33-month sacrifice schedules were added. Otherwise a reconstruction of the liver dose response from the time response would not be possible. There was much greater emphasis on statistical con­ siderations than on toxicological theory in the design of this very large experiment. However, it is remarkable how little attention was paid to Druckrey’s c  t studies on cancer (Druckrey et al., 1963; Druckrey et al., 1967). The interpretation of the study was also driven by statis­ tics and not by the science of toxicology. Revisiting the megamouse study revealed that the occurrence of both bladder cancer and liver cancer was highly consistent with the thesis of Druckrey that c  t  k (Rozman et al., 1996). For the controversy about linearity or nonlinearity of the dose response there is a straightforward biological explanation. 2-AAF is a more potent bladder carcinogen and a less potent liver carcinogen. Therefore, the ED 01 study generated a fairly complete dose response with a shallow and a steep part in terms of bladder cancer with an identifiable threshold. Both the dose-response and the time-response for liver cancer occur to the right of the cor­ responding bladder cancer response, mainly toward the end of the animal’s life span. Therefore, in terms of liver cancer only the shallow part of the dose response is documented by data, because the steep part of it was prevented from developing by the natural life span of the mice. Thus, the low dose linearity is an illusion, which is inconsistent with the theory of toxicology that under conditions of toxico­ kinetic or toxico­dynamic steady state at constant time all dose responses must have also a steep part of their slope. However, if the time-response of a particular effect is trun­ cated by the life span of the species then it is impossible to

Dosage-Response to Radiation  At least one study involving a very large number of animals exposed to varying doses of gamma radiation is available (Ullrich and Storer, 1979a,b,c). Including controls, there were 17,587 mice, of which 15,558 were female. All graphs were on plain graph paper, with inci­ dence of tumors plotted against ­dosage in rads. No attempt was made to explore the logprobit relationship, an omission that, because Gaddum’s famous paper was ­published in 1933, appears unjustified. Because Ullrich and Storer presented raw data, it is possible to explore how their results fit the cumulative lognormal concept. The curves for reticular cell neoplasms in female mice are shown in Fig. 1.9. Here again, the results for thymic lymphoma and for myeloid leukemia are entirely consistent with a linear logprobit relationship that intersects the control level. However, the graph also indi­ cates that, at least within the range of 25–150 rads, increas­ ing doses of radiation caused a decrease in the incidence of reticular cell sarcoma with the result that the total number of reticular cell neoplasms did not begin to increase until the dosage exceeded 100 rads. The authors discussed this phenomenon but failed to consider whether the decrease of one kind of tumor and the increase in other kinds of tumors were independent phenomena or whether the reticular cells that otherwise would become sarcomas were somehow con­ verted by radiation to malignant cells or other configurations that appeared as thymic lymphomas and/or myeloid leuke­ mias. Not shown is a graph for solid tumors in the same mice (Ullrich and Storer, 1979a,b,c). The incidence of ovarian, pituitary, and Harderian gland tumors in excess of controls was consistent with the cumulative lognormal concept. The incidence of lung adenomas was somewhat less at dosage levels of 10–150 rads than it was in the controls, but there was no clear-cut dosage-response relationship such as that for reticular cell sarcoma. Radiation-induced toxicity is a classical, limiting case of toxicology. It is a hit-and-run type poison, whose effects are entirely determined by the dynamics of injury and by expo­ sure frequency. Because the toxicokinetic half-life (resi­ dency time of radiation in an organism) is close to zero and because e0  1, radiation is independent of ­ toxicokinetics

Hayes’ Handbook of Pesticide Toxicology

22

(unless delivered by a carrier such as radon), which simpli­ fies the equation of toxicity to



dT dT dD  ⋅ dE dD dE

Moreover, because radiation-induced injury is extremely rapid it must be recovery from injury that dominates the dynamics of radiation toxicity. In fact, recovery rate constants have been calculated for radiation-induced injury (Sacher et al., 1949; Sacher, 1950 cited in Radiation Biology 1954), although quantitative predictions were deemed to be prob­ lematic. One of the problems was that recovery was nonlinear (Steamer and Christian, 1951), which would have required more data points and curve stripping to separate adaptation and repair rate constants. The other problem was the lack of a clear experimental design to keep all but one timescale con­ stant, when studying a particular effect. The lack of precise diagnostic causes of death introduced further “hidden” vari­ ables in terms of the effect (Steamer, 1951). These are the main reasons for not finding robust quantitative predictions in radiation-induced toxicity. It needs to be reiterated that quan­ titative c  t or c  tx relationships can be seen only under ideal conditions that is lack or very slow reversibility or con­ tinuous exposure to effect and/or regular departures from it (see Section 1.1.3). Discussion  The conclusion that large scale animal studies can neither prove nor disprove the existence of thresholds does not really depend on statistical uncertainty, but on our uncertainty in understanding the basis for any thresholds that exist. We never really accept conclusions that we do not think we understand. As discussed in Section 1.3.7.4, the basis for so-called understanding may vary all the way from very detailed biochemical information (as illustrated by our understanding of the value of vitamins and essential trace elements) to mere economic result (as illustrated by the benefits of food additives for certain livestock). It is not possible to claim the theoretical existence of a threshold on considerations of dose alone because a single molecule in an infinitely large population or in a finite popu­ lation with eternal life could cause an effect. However, it is very straight forward to define a threshold for an effect using both dose and time as variables of toxicity because the maxi­ mum-life-span-dose combination allows calculation of a prac­ tical threshold dose which will have no effect whatsoever in a lifetime in any defined population size (Rozman et al., 1993; Rozman et al., 1996; Rozman and Doull, 1998; 1999; 2000).

1.3.1.4  Kinds of Phenomena Showing a Cumulative Lognormal Form in Their   Dosage-Response Relationships As reviewed in the foregoing sections, the pharmacologi­ cal and lethal effects of compounds on intact organisms

have the form of cumulative lognormal curves, which can be plotted as straight lines following probit conversion. In biology, the log-normal curve is used rarely except in mea­ suring the responses of intact organisms. However, similar curves are obtained when the concentration of a compound is plotted against the inhibition it causes in the activity of an enzyme. The dissociation curves of oxyhemoglobin also have a similar form (Gaddum, 1937). The fact that this form of dosage-response relationship is found in connec­ tion with tissues, enzymes, and macromolecules indicates its fundamental nature. In fact, when the initial concentration (or dosage) of one kind of molecule (e.g., a toxicant) is plotted against the percentage of these molecules reacted with another kind of molecule (e.g., an enzyme or macromolecule) present in excess, the resulting graphs are statistically indistinguish­ able from straight lines within the range of 10–90% reac­ tion. At the extremes, that is, below 10% and above 90%. deviation from linearity is observed. Such effects can be modeled using simple expressions derived from the law of mass action or, more appropriately expressions derived from a consideration of cooperative ligand binding (e.g., the Hill equation). The concept of a threshold can be incorporated into either model. However, a truly realistic model must include kinetic consideration of the rate of inactivation of the poison and the repair of the biochemical lesion. Certainly, this realistic modeling could be complex. However, it might be useful in defining the quantitative aspects of remaining problems. Of course, it is already known that small concen­ trations of a toxicant may be withstood by the intact organ­ ism because it can tolerate some inactivation of enzymes and macromolecules and because critical molecules are replaced in the course of normal repair. Known examples of a chemi­ cal basis for thresholds in dosage-response relationships are discussed in Section 1.3.7.4. The normal distribution is a very deeply rooted phe­ nomenon of nature, which can be found even in nonliving systems. For example, the frequency distribution of the velocity of mercury atoms at 100°C shows a perfect nor­ mal distribution (see Ulich and Jost, 1963). It is interesting that a conversion of the frequency distribution of the veloc­ ities to cumulative frequency would yield a slope similar to the slope of a dose-response under ideal conditions (toxi­ cokinetic or toxicodynamic steady state) particularly if presented on a logarithmic scale. Hayes (1991) was fully aware of the potential complex­ ity of a realistic interpretation of a biochemical lesion when arising as a result of a simple effect (Michaelis-Menten) or of a complex (Hill) effect. However, he did not point out the equivalency of those statements with statements using time as an explicit function of toxicity:

Effect 

MaximumEffect  [ S ] ct Effect  K D  [S ] k



Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

Effect 

MaximumEffect  [ S ]x K D  [S ]

x

Effect 

c  tx k



In the simple case (Michaelis-Menten), both equations depict rectangular hyperbolas, whereas in the cooperative case (Hill), they represent nonrectangular hyperbolas. It is remarkable how much effort went into avoiding time as an explicit function of effects and to express this functional­ ity through substrate concentration and a rate constant (in which time as a variable is concealed) instead of measur­ ing actual time-dependence of effects.

1.3.2  Measurement of Cumulative Effects 1.3.2.1  Early Methods Compounds differ in their tendency to cause cumulative effects when administered repeatedly. The need for a method to express this tendency quantitatively is so obvious that sev­ eral independent attempts have been made to improve on the old method of stating what fraction of a one-dose LD 50 could be tolerated daily. Apparently the first method that involved statistical concepts was that of Lim et al. (1961). It consisted of an increasing dose, two-group (ID2) sched­ ule according to which each drug was started at 9% of the acute LD 50 rate and was increased by a factor of 1.5 every 4 days until the animals died (usually in 3–4 weeks). The starting point and the rate and interval of increasing the dos­ age were chosen in such a way that most tests could be fin­ ished in 24  4 days. The chronic LD 50 was expressed as a percentage of the acute LD 50 for the same compound. Lim et al. (1961) showed that their method distinguished suc­ cessfully (and rapidly) between drugs that were known to be cumulative in effect (chronic LD 50, 5–71%), those that show no important cumulation or tolerance (chronic LD 50, 91–102%), and those that induce tolerance (chronic LD 50, 137–467%). The use of increasing doses (which corresponds to few if any situations in the medical use of drugs or in the exposure of anyone to xenobiotics) apparently was sufficient to discourage adoption of this method. The work of Lim et al. (1961) affords the opportunity to illustrate the power of the Rozman and Doull theory to under­ stand the early search for empirical explanation/exploration of the relationship between dose and chronicity, the latter being a surrogate word for time as the second independent variable of toxicity. It is important to emphasize that accord­ ing to Fig. 1.1 the first and decisive question to be answered is: Where does the rate-determining (limiting) step(s) come from? Dynamics or kinetics? This can be answered unequiv­ ocally if the dynamic (effect) and kinetic(compound) half lives are known because whichever is slower will provide that (those) step(s). The kinetic half life of HpCDD is in the order of 350 days and therefore according to Rozman et al. (2005) this provides the rate-determining step for this ­ compound. Thus, the differential equation on page three applies to this

23

compound. Because of essentially linear accumulation of HpCDD, the acute LD 50 of this compound would have been reached before day 20 after Lim et al. (1961)’s dosing regi­ men. Clearly, he had some compounds with relatively long half-lives (resepine, emetine, digitoxin, bromide) and there­ fore accumulation of compound. However, none had near the very long half-life of HpCDD. His second category is easi­ est to understand. If the acute LD 50 is much less than the (cumulative) (sub) chronic dose then it amounts to giving single daily dose rates with either nearly complete elimina­ tion (kinetic) or recovery (dynamic) every day. In that case the last daily dose rate would have been about 90% of the acute LD 50. Clearly, he had in this category compounds of very short kinetic half lives (acetylsalicylic acid, nalorphine) and therefore no accumulation of compound and for lack of a long dynamic half-life also no accumulation of effect. His last category, depicted as developing tolerance, con­ sists mostly of two types of drugs: receptor-reactive agents (morphine, phenmetraxine, 1-phenylephrine) and enzyme inducers (prochlorperazine, chlorpromazine, phenobarbital, pentobarbital). Tolerance – a form of adaptation – is usually associated with legal or illicit drugs. Development of toler­ ance to the first type of drugs is clearly due to receptor up­regulation, which is a dynamic phenomenon. Since all the above drugs have short kinetic half-lives it is obvious that the rate-determining step (responsible for the up-­regulated recep­ tor steady state) originates from the slow recovery half-life of the receptor up-regulation. The four enzyme inducers all have kinetic half-lives in excess of 1 day, implying that it takes 4 days to reach kinetic steady state and as many to return to base line. Increased synthesis/degradation of enzyme pro­ tein is faster than that, implying that the rate-determining step originates in the kinetics of these drugs, the end result of which is enzyme induction resulting in increased clearance of these drugs. Thus, understanding tolerance in light of the Rozman-Doull theory allows to understand tolerance as orig­ inating either in the dynamics or the kinetics of the underly­ ing adaptation. The first statistical method for measuring cumula­ tive effect that remains in use was developed in the Soviet Union (Kagan, 1964; Kagan and Stankevic, 1964). According to this method, the result is expressed as a cumulation coefficient (Kcum). Whereas a definition of this term (Kcum  LD 50n/LD 501 or Kcum  chronic LD 50/acute LD 50) has been widely circulated in English (Kagan, 1970, 1975), a detailed explanation of the method that would permit any investi­ gator to apply it to his or her own data has not been read­ ily available. Such a detailed explanation did appear as Annex III in a book issued by the All-Soviet Scientific Research Institute of Hygiene and Toxicology of Pesticides, Polymers, and Plastic Materials (Hygiene and Toxicology of Pesticides, Polymers, and Plastic Materials, 1969, here­ after); and the method is discussed in Section II, 2 of that text (pp. 25–29). This book may have escaped the attention

24

of some English-speaking toxicologists. Fortunately, criti­ cal parts of the report were translated (but not published) by the World Health Organization and were made available by Dr. M. Vandekar. According to Dr. Kagan’s method, the chronic LD 50 is determined by plotting (on logprobit paper) each total dosage against the percentage mortality it produced. The LD 50 is then read off the graph from the 50% mortality intercept. The total dosage for each animal is determined by multiplying the number of days it survived by its daily intake of toxicant (mg/kg/day) derived from measured food intake and from the concentration of the toxicant in the feed available to the particular group. Only one point is plotted for each day on which one or more animals die. Thus, for a group in which 10 animals receive a compound at a dos­ age of 1.5 mg/kg/day and in which one animal dies on day 10, two more die on day 12, and so forth, the total dosage plotted for 10% mortality is 15 mg/kg; for 30% mortality, 18 mg/kg; and so forth. The results for each group of ani­ mals are plotted separately, providing a single LD 50. Each LD 50 is divided by the same one-dose LD 50 that has been determined in the usual way in a separate experiment; the quotient is the coefficient of cumulation (Kcum) for the particular group and, therefore, for the particular dietary level of compound. No way of statistically combining the results from different groups has been suggested. Instead, the co­efficients obtained from different groups are evalu­ ated separately in terms of their numerical relationship to the daily dosages that contributed to them. Various methods of using these coefficients to evaluate the cumulative effects of a compound are in the literature (Kagan, 1970, 1975; Hygiene and Toxicology of Pesticides, Polymers, and Plastic Materials, 1969). According to one method, coefficients less than 1.0 signify high cumulation; those greater than 5.0 signify slight cumulation. Kagan’s method was used to determine coefficients of cumulation for warfarin and parathion, using the same raw data used earlier (Hayes, 1967b) to measure the chronicity index for those compounds as described in Section 1.2.2.3. Briefly, the methods agreed in showing that warfarin pro­ duces cumulative effects and parathion does not. However, use of the coefficient of cumulation presents practical diffi­ culties. The exact coefficient obtained for a compound var­ ies with the dosage chosen. The major difference between the method of Kagan and those of Boyd and Hayes is that results from only a single group of animals are considered in a single statistical maneuver. Even when several groups are studied and the different resulting coefficients are plot­ ted to form a curve, there is no statistical integration of the results for the separate groups to generate a single coeffi­ cient. By contrast, each of the other two methods requires (a) the use of several dosage groups and (b) the statistical integration of the results from all of the groups studied. It is concluded that Kagan’s method is not as precise as either of the basically similar methods proposed by Boyd or Hayes.

Hayes’ Handbook of Pesticide Toxicology

Kagan’s method is easily understood in the context of Section 1.1.2 in combination with Fig. 1.1. He plotted the cumulative dose rates (dose) of multiple-dose-rate studies vs. effect on a logprobit paper, which is theoretically legiti­ mate, because the dose is always the sum of all dose rates. Accumulation of effect is either the result of accumulation of toxicant or accumulation of injury, whichever has the longer half-life will dominate the dynamics of injury. If there is no recovery from injury or no elimination of toxicant, then there will be linear accumulation of injury or toxicant according to a triangular geometry after multiple dose rates or accord­ ing to a rectangular geometry after a single loading dose rate followed by maintenance dose rates. Thus the lowest theo­ retical value for Kcum  0.5. Indeed a value of less than 1.0 was considered an indication of accumulation, whereas one greater than 5 was considered an indication of slight accumu­ lation only (Hayes, 1991). The explanation for warfarin caus­ ing and parathion not causing a cumulative type of effect is a direct consequence of an appropriate understanding of the role of time in toxicology. The kinetic half-lives of both war­ farin and parathion are shorter than their dynamic half lives. Therefore, the dynamic half-lives will dominate the actions of both of these compounds. The dynamic half-life of war­ farin is about 1 day (Nagashima et al., 1969). Thus, 90 and 99% of dynamic steady state will be reached after 3.32 and 6.64 dynamic half-lives of continuous exposure, respectively, after which accumulation of effect will occur according to c  t  k with only a 25% difference between Cmax and Cmin assuming two bouts of feeding per day. The recovery half-life of parathion is less than 12 h estimated based on data available for soman (Rozman, 2000a). Thus, steady state for parathion will be reached more rapidly but at an average level less than half of that for warfarin with larger fluctuations between Cmax and Cmin. Depending on the dose selection, the cumulative effect of warfarin may be above the 90-day c  t threshold whereas that of parathion may be below it.

1.3.2.2  C/A Index The next statistical method to be introduced was that of Boyd et al. (1966), who suggested that the comparison be made at the LD 50 levels for both the acute and the sub­ acute tests. Specifically, he proposed that a one-tenth life span (0.1L) chronic/acute LD 50 (0.1L) index [C/A LD 50 (0.1L) index] be calculated by expressing the multiple-dose LD 50 as a percentage of the acute LD 50. (Both kinds of LD 50 values involved stomach tube administration, but the acute dose was given to nonfasted animals.) The C/A index for sodium chloride was found to be 72, indicating that 100 daily doses of table salt each at a rate 72% of the acute LD 50 would kill half of a very large population of rats. The C/A index, unlike Kagan’s cumulative-dose-rate (  dose) vs. single-dose-rate coefficient, relates the single dose rate of a multiple-dose-rate study to a single-dose-rate experiment (chronic LD 50/ acute LD 50) and expresses this

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

ratio as a percentage. Hayes (1991) used the reciprocal of that ratio multiplied by 100 as an index to characterize the cumulative nature of an effect. He was also aware of the dif­ ference between accumulation of compound and accumula­ tion of injury, and that his index did not distinguish between the two. Comparing multiple dose rates (daily doses or dos­ ages) with a single dose rate (single dose) is problematic because it confuses the issue that no matter what happens to the toxicant, the organism was still exposed to the sum of the dose rates, which is the dose, just as the organism is exposed to the total dose after administration of a single dose rate. The lack of conceptualization of subsequent events which occur on different timescales led to a great deal of confusion in toxicology. As eminent a scholar as Druckrey held the erroneous view about reinforcing effects of small dose rates, because he compared single daily doses (dose rates) instead of doses (sum of dose rates) in diethylnitrosamine-induced

25

cancer. The dose rate (daily dose) causing 100% cancer in about 60 days was 200 times higher than that causing cancer in about 900 days. In fact, the doses were only different by a factor of 12, which represents the specific time response (Fig. 1.2) at constant dose rate (steady state). Both the C/A index and the index of Hayes are useful as indicators of chronicity. Nevertheless, their lack of dis­ tinction between the different timescales involved conveys incomplete and often erroneous messages. High chronicity index for mirex is indicative of strong cumulative toxicity. However, the more important information that the cumula­ tive AUCs after a single dose rate (1-dose LD 50) of mirex are virtually the same as after 90 fractionated dose rates (90-dose LD 50), indicating a slight (40%) adaptation to mirex, is not revealed. Similarly missing is the important information that the toxicity of mirex is largely dominated by its kinetic half-life (ATSDR, 1995a) of about 350 days

Table 1.2  Absolute and Relative, Acute and Subacute Oral Toxicity of Certain Pesticides and Drugsa Compound

Species

Sex

1-Dose LD 50 (mg/kg)

90-Dose LD 50 (mg/kg/day)

Chronicity index

Mirex

Rat

F

365

6.0

60.8

Warfarin

Rat

M

1.6

0.077

20.8

Metepa

Rat

M

136

7.5

18.1

Dieldrin

Rat

M

102

8.2

12.8

Atropine

Rabbit

M

588b

78b

7.5

Apholate

Rat

M

98

17

5.8

Paraquat

Rat

F

110

20.5

5.4

DDT

Rat

M

250

46.0

5.4

Benzylpenicillin

Rat

M

6700c

4140c

1.6

M

d

e

1.4

Sodium chloride

Rat

3750

Caffeine

Rat

F

192

Parathion

Rat

F

3.6

f

2690 150

g

1.3

3.1

1.16

3.5

1.03

Azinphosmethyl

Rat

F

11.0

10.5

1.05

EPN

Rat

F

7.7

12.0

0.64

Dichlorvos

Rat

F

56

70

Potassium cyanide a

Rat

M

10

250

0.08 h

0.04

From Hayes (1967b) or later U.S. Public Health Service data, except as noted, by permission of Academic Press. The compounds are listed in approximate order by decreasing chronicity index. b Boyd and Boyd (1962) (100-intramuscular-dose test). c Boyd and Selby (1962 (100-dose test). d Boyd and Shanas (1963). e Boyd et al. (1966) (100-dose test). f Boyd (1959). g Boyd et al. (1965) (100-dose test). h No mortality occurred at 250 mg/kg/day, the highest dosage administered.

Hayes’ Handbook of Pesticide Toxicology

26

(in rats). The toxicity of warfarin represents the opposite end of the spectrum because it is determined by the dynamics of the effect as discussed earlier. A chronicity index of 20.8 is indicative of cumulative toxicity, but it is misleading in that it suggests the toxicity is less cumulative than that of mirex. In fact, considering that no rat died related to administration of warfarin in the 90-dose-rate study after day 25 shows that the chronic dose (sum of dose rates) was a mere 1.2 higher than the acute dose (one dose rate) indicating nearly perfect c  t  k, implying chronicity similar to mirex.

1.3.2.3  Chronicity Index The chronicity factor introduced independently by Hayes (1967b), is expressed as a quotient rather than a percent­ age. However, this factor is really an index and ought to be designated as such in the future. Excluding differences in the procedures for measuring the LD 50 values, the chronicity index for a compound is the reciprocal of its C/A LD 50 (0.1L) index expressed as a fraction instead of as a percentage. That is, Chronicity index 

100 C/A LD50(0.1L )index

For example, the C/A LD 50 (0.1L) index for sodium chloride (71.7) would correspond to a chronicity index of 1.395. Because each chronicity index is a ratio, these indices may be used to compare the tendency of different com­ pounds to have cumulative effects without reference to their absolute toxicities. This index is determined on the basis of an observed effect. No distinction is made between effects that depend in part on cumulation of the toxicant (e.g., lead) and those that do not (e.g., alcohol). The chronicity index for each compound is obtained by dividing its 1-dose LD 50 (expressed as milligrams per kilogram) by its 90-dose LD 50 (expressed as milligrams per kilogram per day). The resulting number is large (2.0 or more) for compounds that are relatively cumulative in their effects and small (less than 2.0) for compounds that show little cumulative effect. The index of 2.0 is recognized as an arbitrary dividing point, but it appears supported by the limited data available and is also plausible on theoreti­ cal grounds. In any event, if a compound were absolutely cumulative (in the sense that 1/90 of the 1-dose LD 50 was exactly the 90-dose LD 50), the chronicity index would be 90. A chronicity index of 1.0 associated with oral intake indicates that daily ingestion of the 1-dose LD 50 mixed into the regular diet leads to death of half of a very large population so exposed for 90 days, which is very difficult to verify experimentally, but sodium chloride gets as close to it as experimentally possible. Table 1.2 summarizes the 1-dose and 90-dose LD 50 val­ ues and also the chronicity indices for warfarin and several

other compounds. The marked cumulative effect of war­farin and the chemosterilants; the small magnitude of such an effect of table salt, caffeine, and some organic phosphorus compounds; and the essential lack of cumulative effect of potassium cyanide are recorded. The 90-dose LD 50 of war­ farin was only about 1/20 of the 1-dose LD 50, indicating a chronicity index of about 20 for that compound. It required daily ingestion of approximately a 1-dose LD 50 of several organic phosphorus insecticides to kill half of the test ani­ mals in 90 days, indicating a chronicity index of approxi­ mately 1 in each instance. Rats tolerated daily 25 1-dose LD 50s of potassium cyanide mixed with their regular food with no mortality, indicating a chronicity factor of less than 0.04. This tolerance for organic phosphorus compounds and cyanide undoubtedly indicates the ability of the body, and especially the liver, to detoxify moderate dosages of these materials provided there is time in which to accomplish the task. The chronicity index permits comparison of the effects of different classes of compounds of the same class. Whether these smaller intraclass distinctions are really significant or whether they are outweighed by differences caused by spe­ cies or other factors must be determined by future experi­ ence. It is certainly to be hoped that increasing use will be made of 90-dose LD 50 and ED 50 values and of the chro­ nicity index in order that the study of long-term toxicity may be made more quantitative. The chronicity index is a measure of cumulative effects. A concentration index has been proposed as a measure of cumulative storage. The effect of a compound cannot be less than that determined by its storage in the body, espe­ cially its presence (storage) in sensitive tissues. In this sense, a compound that has a high concentration index will tend to have a high chronicity index. However, some com­ pounds are highly cumulative in their effects even though they show a minimal tendency to storage. Thus, the two indices do not vary in a parallel fashion. It is generally agreed that what has been called biologi­ cal magnification is the basis for the injury caused by DDT and a few other compounds to certain large, predatory forms of wildlife. Biological magnification occurs in situ­ ations in which a compound shows a high concentration index in each successive species in a food chain. This section demonstrates that Hayes (1991) was fully aware of the importance of time in the manifestation of toxicity without generalizing time as an equivalent and fully quantifiable variable of toxicity along with the dose. Perhaps for this reason, he made no reference to measur­ ing time accurately in toxicological experiments. There are additional issues to be considered when viewing time as a variable of toxicity: the timescale on which the effect is occurring and the frequency of observation, which are related to each other as well as to the half-lives of com­ pound or effect and exposure frequency. A clear distinction must also be made whether dose-time or effect-time rela­ tionships are being considered because the former requires

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

the study to be conducted at constant effect, whereas the latter necessitates an experiment at constant dose (steady state). Routine daily observation of experimental animals in chronic experiments arises out of practicality and is (e.g., cancer studies) without scientific rationale. In fact, in two-year or longer lasting cancer studies weekly observa­ tion would yield satisfactory time resolution of the cancer latency period (but not for harvesting tissues). However, the daily observation of animals in acute experiments often provides worthless information on that timescale if all the animals die within 2–3 days or even sooner. Automated cameras could provide hourly or continuous monitoring, which would result in the necessary accuracy for quan­ titative time relationships. Toxicant-induced reduced sea urchin sperm motility occurring on a timescale of hours is only meaningfully measured on a timescale of minutes and nobody in his right mind would want to study pungency on a timescale of minutes when it requires time resolution on a scale of seconds. This all sounds simple and straight­ forward, yet cookbook-type toxicology is devoid of these simple considerations. The timescale on which an effect is occurring is important for several reasons. The length of the observation period for any experiment should be the time required by a LOAEL to cause the effect. The relationship between the time to an effect and the dynamic or kinetic half-life of an effect are other critical variables which often confuse toxicological experiments, because the respective rate-determining time ratios will introduce different ratios of intoxication/recovery as discussed in Section 1.3.1.1 with dieldrin and toxaphene as examples. It must be recog­ nized that our understanding of time is perhaps more lim­ ited than mankind’s understanding of matter was during the era of Paracelsus and yet toxicology is one of the few fields that can open the gate to the structure of time.

27

level usually suggests that detoxification and excretion of the toxicant are inefficient, but sometimes means that the toxicant produces some anatomical or biochemical lesion that recovers slowly or not at all. Time relationships in toxicity often can be expressed best by recording the range and mean of time required to produce an observed effect. However, a more elaborate statistical treatment sometimes is indicated. An ET 50 or LT 50 gives a more precise estimate of time to be anticipated in repeated tests than can be expressed by a simple average. The log­ time-logdose curve (Section 1.3.3.2) has also considerable theoretical interest and, in some instances, may be used to predict the proper dosages to be used in long-term studies. In this paragraph Hayes truly anticipated developments in theoretical toxicology that occurred after his death.

1.3.3.1  ET 50 or LT 50 An ET 50 is a statistical estimate of the interval or time from dosage to a specific all-or-none response of 50% of the organisms in a very large population subjected to a toxicant under specified conditions. As used here, an all-or-none effect may be a specified level of a quantitative response: for example, time of appearance of the first tumor or time at which the systolic blood pressure reaches 150 mm Hg. An LT 50 is a special case of an ET 50 in which the effect reported is death. ET 50 and LT 50 values are determined by relating the cumulative percentage effect. In practice, the calculation (Bliss, 1937) or graphic solution (Litchfield, 1949) is carried out in a manner essentially identical to those used for ED 50 and LD 50 values.

99.9 99 95 90

The report of essentially every toxicity test should include information on the time relationships of the effects observed. It is important whether a single dose produces its effect soon after dosing or only after hours or days and whether the effect is brief or prolonged. From a practical standpoint, it is important to know whether a patient, who is mildly sick an hour after overexposure to a toxicant, is really “over the worst of it” or likely to slip at any moment into a critical condition. From a theoretical standpoint, rapid onset of ill­ ness following dosage of experimental animals at or below the LD 50 level suggests that the toxicant is absorbed rap­ idly and acts directly. Rapid recovery following dosing at a substantial rate suggests that the toxicant is excreted or detoxified rapidly. On the other hand, slow onset of illness following dosing at almost any level suggests that the toxi­ cant is absorbed slowly or must be metabolized before it can act. Prolonged illness following dosage at or below the LD 50

Mortality (%)

1.3.3  Time Relationships

50

10 5 1 0.1 0.01

1

3

5 10 30 50 Survival time (days)

100

Figure 1.10  LT 50 curves for male Sherman strain white rats adminis­ tered repeated dosages of warfarin. Dosages (mg/kg/day) were as follows: (O) 0.04; ( ) 0.08; () 0.16; () 0.32; () 0.64; () 1.28. Points adjust according to the method of Litchfield and Wilcoxon (1949) are distin­ guished by a superimposed O.

•

Hayes’ Handbook of Pesticide Toxicology

28

(supraeffective) doses because sometimes the mode of action changes. However, if that is not the case and the mode of action remains the same then there is a highly orderly decline in the slope of the time response regarding mortality (Rozman, 1999) or other effects (Gardner et al., 1977). Figure 1.10 is plotted on a double logarithmic scale which is not as sensitive to deviations from a straight line as is the single logarithmic plot applied by Gardner et al. (1977) and Rozman (2000a). Single logarithmic plots of time-responses yield S-shaped curves similar to doseresponses in log(dose) vs. effect plots, again representing the inverse function of the theoretical plot represented by log(effect) vs. time (arithmetic).

1.3.3.2  Logtime-Logdosage Curve Compounds may show one or more of the following inter­ relations between dosage and time to response, regardless of whether there are one or more doses: (a) a uniform delay between the first dose and the response; (b) a prolongation of the interval that is inversely related to dosage; and (c) a complete absence of detectable effect at low dosages and, therefore, an interval that exceeds the period of observa­ tion (which may be the lifetime of the subjects). Delayed Toxicity  As is well known, there is a lag in the appearance of a detectable effect of some compounds. In other words, there is an inherent delay in their action, which is not accounted for by the time necessary for their absorption

50 30 10 Daily dose (mg)

There is one striking difference in the form of these sta­ tistics for dosage and time. In considering dosage per se, the time of response is ignored completely. A series of tests involving several dosage levels of a compound results in a single dosage-response curve. On the other hand, in con­ sidering time of response, dosage cannot be ignored, and a series of tests involving several dosage levels results in a series of separate curves of different slope. A sufficiently low dosage of any compound will generate a curve coin­ ciding with the baseline, indicating that no animals were affected. The critical range will generate a series of curves such that both the slope and the magnitude of the ET 50 are inversely proportional to dosage. In general, progressive increase of the dosage beyond that necessary to affect all animals will cause progressively less and less change in the slope and position of the ET 50 curves. However, in some instances, progressive increase in dosage beyond that neces­ sary to kill all animals will cause a relatively sudden shift of the very-high-dosage ET 50 curves to the left accom­ panied by an unpredictable change in their slopes. Such a change indicates that a different mode of action has begun. Any dosage above that necessary to kill all organisms in a population is a supralethal dosage, but the term is used most often in connection with dosages that involve some difference in mode of action. Examples may be found most commonly in the toxicology of compounds of which the ordinary effects are delayed. Such compounds are discussed further in Section 1.3.3.2. Except for the phenomenon of changed mode of action, the points discussed in the last paragraph are illustrated by Fig. 1.10 which shows LT 50 curves resulting from differ­ ent dosage levels of warfarin administered in connection with a 90-dose LD 50 study. (Similar LT 50 curves were obtained in connection with a 1-dose LD 50 study.) It may be seen from Fig. 1.10 that, in practice, the progression of curves from right to left is not always completely orderly. The curves at the right tend to be horizontal or incomplete (indicated by dashed lines) because only a portion of the animals in these groups die. The curves at the left tend to approach the vertical, but there is some irregularity, caused no doubt by individual differences and the fact that only a limited number of animals are used in each group. (The data on which Fig. 1.10 was based were used in a differ­ ent form in connection with the corresponding dosage lev­ els in Fig. 1.12 (see Section 1.3.3.2). A comparison of the two figures shows the value of the two kinds of graphs for illustrating different aspects of the same results.) It is amazing how clearly Hayes (1991) saw the problem of time being ignored when “considering dosage per se.” Section 1.1.4 deals with this problem by defining mathematically that the toxic action of chemicals consists of a specific effect at constant dose plus a specific effect at constant time. He was also keenly aware of the fact that the time response curves do not progress in an entirely orderly manner at supralethal

30

15 30

5 3

145 70

70 169

1 5 148

3

158

1 0.5 0.3 10

100

1000

10,000

Time for appearance of first tumor (days) Figure 1.11  Response of rats to graded daily of 4-dimethylaminoazo­ benzene (4-DAB) administered orally. No liver tumor was obtained with the two lower doses employed, indicated by arrows. The number associ­ ated with each point or arrow refers to the number of animals tested at that dosage. The dotted line represents a life span of 2.74 years. From data of Druckrey (1959).

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

and distribution to the target organ. The inherent delay is not fully overcome by substituting larger doses or by using intra­ venous injection. Examples are offered by (a) carcinogens, for which an induction period apparently always is required, (b) certain organic phosphorus compounds that produce paralysis in humans or chickens, but only following a delay of about 10 days, and (c) the coumarin-derived anticoagu­ lants, which inhibit prothrombin formation but have little or no initial effect unless the dosage is massive and a different toxic mechanism is involved. Some other compounds, nota­ bly alkylating agents (Hayes, 1964), produce delayed effects but generally produce some illness promptly after a dose at or even below the LD 50 level. A compound cannot produce a delayed effect unless it or its metabolites or a direct or indirect pharmacological action persists until the clinical effect appears. This persis­ tence of a compound or its action is the essence of which cumulative effects are made. On the contrary, the couma­ rin-derived anticoagulants exhibit the delay but do not pro­ duce a truly chronic disease. Whereas the biochemical basis for the delayed action of coumarin-related anticoagulants 100

Dosage (mg/kg/day)

50 30 20 10 1

5 3 2 1 0.5 0.3 0.2

2 3

0.1

LD 50

0.05 0.03 0.02 0.01

LD 10 LD 1 LD 0.1 LD 0.01

1

3 5 10

30 50 100

300

1000

Time (days) Figure 1.12  Relationship between dosage of warfarin and time of death in rats. Short horizontal line indicates the time of death of the first animal in a group to die, and a solid point indicates the geometric mean time to death for a group in which all the animals died. In those groups in which some animals did not die, their survival is indicated by an arrow, and the best estimated geometric time to death is indicated by an open circle. The tip of each arrowhead indicates the end of the dosage period. Because it could not be assumed without produce that the survivors would live a nor­ mal life span, it was empirically assumed for purposes of calculation that the survivors died on the last day of the test. Thus, the true position of these estimated values always lies to the right of the open circle. Note that no rats died when dosed for 300 days at a rate of 0.02 mg/kg/day. The graph also shows the 90-dose LD 50 and some other dosage response values for 90 doses of warfarin. Dotted vertical line represents average life span. Slightly modified from Hayes (1967b), by permission of Academic Press.

29

is understood clearly, the reasons for the delays associated with other agents are obscure, probably because we do not yet know the biochemical lesions involved. It is conven­ tional to explain the delay associated with carcinogenesis as the result of a “multistage process” but this is more a phrase than an explanation. The term “delayed toxicity” ought to be restricted to delay in onset of clinical effects following the absorption of an adequate dosage. The term ought not to be used to refer to (a) a delay in onset that depends on the time neces­ sary for the accumulation of an adequate total dosage from relatively small repeated exposures or (b) the progression of disease, including any that is the result of scarring or some other morphological or biochemical effect that is an inherent part of the toxic injury. Although the necessary delay in the onset of the effect of some compounds is well known, it has not been custom­ ary to represent it graphically. By contrast, curves relating the increasing interval from the first and sometimes only dose until the appearance of a selected effect were used at least as early as 1937 and were based on data published as early as 1908 (Clark, 1937). Clark showed that, for a certain range of dosages characteristic of each test system, there often is a linear relationship when the logarithm of time to response is plotted against the logarithm of dosage producing the response. Both the graphic and the mathe­ matical features of the relationship were thoroughly inves­ tigated by Bliss (1940). Figure 1.11 indicates the form of a typical curve. Recognition of the relationship apparently has not been general, with the result that it has been redis­ covered from time to time. It is a general principle of toxicology that any compound may be tolerated without injury provided the dosage is suffi­ ciently small. It has not been customary to represent this rela­ tionship graphically. As discussed later, such ­representation is desirable, for it reveals what may be basic differences in the behavior of different compounds. Shape of the Complete Logtime-Logdose Curve  Summar­izing the last several paragraphs, it is evident that a complete logtime-logdose curve would have three segments: 1. The first segment represents the minimal time neces­ sary to produce an effect even with dosages larger than the minimal one required. 2. The second segment represents the increasing times necessary to produce an effect with successively smaller dosages. 3. The third segment indicates a dosage a little below which the effect is not produced, no matter how long dosing may be continued. Such a curve based on a study of warfarin is shown in Fig. 1.12. The three segments are well shown. The second

Hayes’ Handbook of Pesticide Toxicology

30

segment was established by dosages of 0.16, 0.32, 0.64, and 1.28 mg/kg/day, and in each instance the time necessary for half of the animals to die was 10 days or less. (The points in Fig. 1.12 represent geometric means as explained in the legend. LD 50 values could have been used and would have given essentially identical results. The choice was based on convenience, especially in connection with groups in which only a few animals died.) As may be seen, the third segment in Fig. 1.12 has been drawn out horizontally at the level of the 90-dose LD 50 value for warfarin as determined by the original, detailed form of the 90-dose curve in Fig. 1.6. The corresponding LD 10, LD 1, LD 0.1, and LD 0.01 have been indicated also. Because the lowest dosage tested (0.02 mg/kg/day) lies between the values calculated to be the LD 0.1 and LD 0.01 levels, there is little wonder no effect was observed among a group of only 10 animals. It appears that a few compounds (e.g., warfarin in Fig. 1.12) exhibit all three segments of the theoretical curve, some compounds (e.g., 4-DAB in Fig. 1.11) exhibit the first and second segments only, and most compounds exhibit the second and third segments only. Perhaps some compounds exhibit the second segment only, but no illus­ tration is available. It is impossible to make a more exact statement at this stage because there has been so little study of comparative, quantitative toxicology. In fact, it is not established that all compounds exhibit a typical sec­ ond segment of the theoretical curve, although this appears likely. Curved second segments shown by some authors (Clark, 1937) may, in fact, represent a transition between second and third segments. The presence of a delayed reaction following large dos­ ages does not exclude the possibility that small dosages of the same compound may be tolerated. Figure 1.12 offers an illustration of this kind of tolerance in rats fed warfa­ rin. Another example is offered by the work of Siegel et al. (1965), who showed that a mixture of tricresyl and other triaryl phosphates, which produce paralysis of chickens and rabbits after only 20 days of high-level exposure, was tolerated by both species for as long as 90 days when given at lower dosage levels. Of course, it is easy to demonstrate tolerance for small dosages of most compounds, albeit they do not elicit a significant latent period when absorbed at high dosage levels. In all instances studied so far, carcinogenesis is asso­ ciated with lack of a third segment in the logtime-logdose curve. In connection with toxicity generally, lack of this segment is exceptional. The presence of the third segment strongly implies the existence of a threshold at a level only a little less than the level of the segment itself. It is not cer­ tain how the absence of a third segment ought to be inter­ preted, but there is no evidence to exclude the possibility that a threshold exists here also at a dosage level just below that required to produce the smallest statistically signifi­ cant increase in the incidence of tumors above control

Substances

First used by

c  t

Ethylbromoacetate

France

3000 and less

Chloroacetone

France

3000

Xylylbromide

Germany

6000

Chlorine

Germany

7500

Perchlormethyl mercaptan

France

3000 and less

Hydrocyanic acid*

France

1000

Phosgene

France

450

Methylchloroformate

Germany

500

*The value of (c  t) for hydrogen cyanide depends on its concentration. The value given refers to the concentration of 0.5% obtainable in the field. The values are much higher with smaller concentrations. (Reprinted from Haber (1924), by permission of Springer-Verlag.)

levels in animals that survive as long as any of their spe­ cies. Regardless of the logtime-logdosage-response, the existence of a threshold should be demonstrated biochemi­ cally, as discussed in Section 1.3.7.4. Use of the Logtime-Logdosage Curve for Prediction Aside from its basic interest for toxicology, the logtimelogdosage curve may be used in connection with a brief test to predict appropriate dosage to use in long-term studies. Reference to Fig. 1.11 shows that only about 90 days of testing of 4-DAB in rats at dosages of 10, 18, and 32 mg/rat/day would have been sufficient to predict that a dosage of about 0.9 mg/rat/day produces an effect within the lifetime of that species. On the contrary, use of the same technique in connection with warfarin (Fig. 1.12) predicted a limiting value which, however, did not correspond closely with the value actually observed in a long-term experi­ ment. By extrapolation of the second segment of the curve (as shown in part by the dashed line in Fig. 1.12), one would predict that a dosage of approximately 0.002 mg/kg/day would kill half of a sufficiently large group of similar ani­ mals within 90 days. This prediction for warfarin is seri­ ously inaccurate when compared with the 90-dose LD 50 of 0.077 mg/kg/day based on all the dosage levels tested, including a dosage of 0.020 mg/kg/day, which was toler­ ated for 300 days without any mortality. The fact that the value predicted may be only limiting and may not correspond closely with observed long-term toxicity does not make the test useless. The test does have the advantage of relative brevity. It is better to know a lim­ iting value than to have no valid guide for choosing dos­ ages for long-term study. The logtime-logdosage curve also may be used in the bioassay of bacterial and other toxins. What may have been

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

the first effort along this line (Boroff and Fitzgerald, 1958) confirmed the linear log-log relationship demonstrated ear­ lier for other substances and showed that, by the intravenous injection of relatively large doses, a test could be completed in less than two hours instead of the four days required for the conventional test for toxins. It was shown later that, by using dilutions that had been tested in the conventional way, it was possible to construct a standard curve relating the log­ arithm of the mean survival time in minutes to the logarithm of concentration or logarithm of the number of lethal doses per unit volume, thus providing a prompt measurement expressed in the desired unit (Boroff and Fleck, 1966). Haber’s Rule Apparently the only statement Haber made of what has come to be called his rule is contained in a footnote to the last of a series of five lectures this Nobel Prize-winning chemist made in the years 1920–1923 (Haber, 1924). This particular lecture concerned the history of gas warfare, and all the toxicological considerations were in this context. This means, among other things, that only brief exposures were considered. At that time, no chemical was known that would not drift away or be diluted to a harmless ­concentration soon after its release. The complete footnote may be translated as follows: A simple and practical measure for toxicity can be obtained that suffices for all practical purposes. For each war gas, the amount (c) present in one cubic meter of air is expressed in milligrams and multiplied by the time (t) in minutes necessary for the experimental animal inhaling this air to obtain a lethal effect. The smaller this product (c  t) is, the greater is the toxicity of the war gas. A few values obtained during the war are given in the table. More detailed information can be found in the medical literature. The values were all obtained by using cats as experimental animals. The chemicals are listed according to the order of their introduction as war gases. It may be noted that the footnote implies but does not state what is now called Haber’s rule for equitoxic doses, especially fatal doses, namely,

ct  k

where c is concentration, t is time, and k is a constant char­ acteristic of a particular compound. Actually, the concept was not original with Haber. Apparently it was stated first by Warren (1900) in connec­ tion with his studies of the effect of different concentra­ tions of sodium chloride on Daphnia magna. Warren stated the relationship as

T ( X  8)  constant

where T is the time of killing and X is the strength of salt solution. The value 8 was an observed constant concentra­ tion below which the relationship did not hold and survival of the animals was influenced little or not at all by the salt.

31

Table 1.3  Dosage Relationship for the Second Segment of the Curve in Fig. 1.12 Daily dosage (mg/kg)

Total dosage (mg/kg)

1.28

7.17

0.64

4.28

0.32

2.69

0.16

2.03

Thus, Warren recognized that the relationship was true only within certain limits of time and concentration. It is clear from his note on hydrocyanic acid that Haber, too, was fully aware of the limitation of the constant relationship. Bliss (1940) reviewed some earlier papers on the sub­ ject and presented an elaborate mathematical analysis of the relationship between exposure time and concentration. Restatement of Haber’s Rule  Further study has empha­ sized the limited applicability of the rule. Recognition of this limitation has led to a restatement of the relationship as where



[(C ⋅ Vm )  De ] ⋅ t ⋅ R D w

D  dosage (mg/kg) received during time t C  concentration of toxicant (mg/m3) Vm  minute volume rate of respiration (m3/min) De  detoxification rate (mg/mm) t  time (min) of exposure w  body weight (kg) R  retention coefficient expressed as a decimal fraction This equation shows that a sufficiently high rate of detoxi­ fication would negate prolonged exposure to a sufficiently low concentration. It thus expresses quantitatively the limitation on the rule when applied to easily detoxified materials like hydrocyanic acid. It will be seen that in this equation dosage (D) is not necessarily a constant for all combinations of concentration and time that produce the same effect, because the detoxification rate and perhaps the retention coefficient may vary with dosage. David et al. (1981) evaluated the role of time and con­ centration on carbon tetrachloride toxicity in rats. Using hepatotoxicity as a marker and varying concentrations and time products as exposure, these authors concluded that the severity of liver lesions was more influenced by the con­ centration of carbon tetrachloride in the inhaled air than by the product of concentration and time. The limiting vari­ able was the length of time required for tissues to acquire critical concentrations of the toxin.

Hayes’ Handbook of Pesticide Toxicology

32

Relation of Haber’s Rule to the Logtime-Logdosage Curve  If one considers the relationship c  t  k which constitutes Haber’s rule, it is clear that it represents a spe­ cial case of the second segment of the logtime-logdosage curve (when the dosage is expressed as concentration). When plotted on log-log paper, all solutions of the equation lie on a straight line passing through two points, namely,



c1 tk ck t1



Furthermore, on the same set of coordinates, all solu­ tions of all other equations of the same form will lie ­parallel to the first but pass through k1, k2 and so forth, instead of k. The slope of these lines is algebraically the same and is 1 on ordinary log-log paper. Some logtime-logdosage curves based on observed data have a slope statistically indistinguishable from that deter­ mined by Haber’s rule. Figure 1.11 shows an example. However, other real curves show greater or lesser slopes (Bliss, 1940; Clark, 1937; Druckrey, 1943, 1967; Scholz, 1965). Examination of Fig. 1.12 shows that the second ­segment has a downward inclination steeper than 45°. In other words, within the range of dosage from 1.28 to 0.16 mg/ kg/day, the smaller dosages are progressively more effective than would be predicted by Haber’s rule. That is, progres­ sively less total dosage is required as shown in Table 1.3. In this instance, the reason is that warfarin does not depress appetite so that, at 1.28 and 0.64 mg/kg/day, the rats con­ sumed relatively large total dosages of this slow-acting compound before they had time to die. On the contrary, at an intake of 0.02 mg/kg/day, no injury occurred even though a total dosage slightly over 5.0 mg/kg was taken in during a period of over 300 days. With the exception of the mat­ ter just discussed, the significance of slopes greater than or less than 1 is not clear; in fact, both have been observed for warfarin under different conditions. In any event, there is no relationship between the slope of the second segment of the curve and the occurrence of a third segment. Discussion  There is no meaningful relationship between Haber’s rule and time-weighted averages for occupa­ tional exposure although they apparently have been con­ fused. Both concepts involve time and concentration, but Haber’s rule is an equation describing a principle of toxi­ cology whereas a time-weighted average is a standard set to prevent overexposure of workers or others. As already stated, Haber’s rule is a special case of the second segment of the logtime-logdosage curve. Each time-weighted aver­ age defines a level of exposure to a particular compound intended, often on the basis of extensive experience, to be tolerated without injury for a lifetime; this level of permis­ sible exposure lies below and parallel to the third segment of the logtime-logdosage curve. The distance between the

second and the third segment of the curve is a safety factor the magnitude of which will vary with the compound. Investigating dose-time relationships requires a very accurate definition of effect because the experiments have to be conducted under isoeffective conditions. For example, it is not enough to state that the effect of interest is lung can­ cer. It needs to be specified whether it is time to first lung cancer or time to 50% lung cancer. Further specifications are needed regarding the severity of the effect: time to histo­ logically identifiable cancer obviously should not be lumped with time to death caused by lung cancer. Even though these considerations may appear trivial, they are pivotal and unfortunately routinely ignored in the design and interpreta­ tion of toxicological experiments. The log(dose) vs log(time) plot has its theoretical justi­ fication in Haber’s c  t  k concept, the logarithmic form of which is

log c   log t  log k

which is the equation of a straight line (Fig. 1.11). The arith­ metic form (c  t  k) provides a rectangular hyperbola with the limits set by the minimum lag time of the effect and max­ imum life span of the species studied (Fig. 1.12). It is also important to recognize that theoretically all effects have a lag period between dosing and effect. Inhalation anesthesia has a very short lag period whereas cancer has a very long lag time called latency. The minimum lag period is a characteristic of the effect and therefore not subject to change. For example, HpCDD-induced delayed acute toxicity cannot cause lethal­ ity in less than 8 days even if supralethal doses are given to rats (Rozman, 1999). This lag period can have either kinetic/ dynamic (delayed absorption or slow accumulation of com­ pound) or dynamic reasons (delayed time to effect or slow accumulation of effect). Ingestion of acutely nontoxic dose rates of lead can cause a fulminant episode of toxicity once the reserve (storage) capacity of the organism has been exhausted because of accumulation of lead. A single high dose of diethylnitrosamine will cause cancer in 100% of the animals because of the persistence and hence slow recovery of the DNA damage. Section 1.0 deals in great detail with all other issues involving time that were raised by Hayes (1991). Even more information may be found on the role of time in risk assessment (Rozman, 2000a) and the use of the c  t concept in the context of establishing occupational exposure levels. Thus, we do not share Hayes (1991) view that there is no meaningful relationship between Haber’s rule and time weighted averages in occupational exposure. The conditions are clearly outlined in Section 1.1, which define the condi­ tions for Haber’s rule of inhalation toxicology to become a fundamental law of toxicology. There are no exceptions to this law that we have found and therefore we suggest that any experiment that appears not to obey this law is incom­ pletely designed or controlled. However, for Haber’s rule to become manifest requires continuous exposure. Occupational

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

33

e­ xposure is discontinuous (8 h/day, 5 days/week for 45 years), which allows for recovery or elimination 16 h between work­ days and 66 h on weekends. Therefore, there is a deviation from the c  t  k or c  t  k  W concept in the form c  tx  k for isoeffective or c  tx  k  W for isodosic or isotemporal exposures in an occupational setting which requires the transformation of c  tx back to c  t in order to establish meaningful exposure limits. This issue has been discussed and clarified by Doull and Rozman (2000).

responses of organisms are graded in character and need to be so reported. Such reports may take many forms, includ­ ing tables and line and bar graphs, and represent the whole gamut of pharmacological effects. In some instances, the data may be treated mathematically. Some examples are given in the section on storage. However, graded responses do not lend themselves to neat quantitative tabulation such as may be applied, for example, to the LD 50 values of a series of compounds.

1.3.4  Problem of Measuring Effect of Dispersed Toxicants

1.3.6  Dosage at the Tissue Level

Although it is desirable to express dosage in terms of weight of toxicant per weight of organism, this is difficult if the dose is absorbed from air by the lungs or trachea of land animals or from water by the gills of aquatic animals. Under these circumstances, it may be convenient and even necessary to consider toxicity in terms of concentration of the toxicant in the medium. If there is continuous exposure to constant concentrations, the data are expressed in the form of EC 50 values as explained later. If time is also a variable, the time results may be presented in terms of a logtime-logconcentra­ tion curve as discussed in connection with Haber’s rule. When possible, dosage in terms of milligrams per kilo­ gram should be calculated from the concentration of toxi­ cant, the respiratory volume, and the proportion of toxicant retained. The result of this calculation offers one impor­ tant way of comparing toxicity by the respiratory route with that by other routes. Another approach is to measure plasma or other tissue levels of toxicant following expo­ sure by different routes (Section 1.3.6).

1.3.4.1  EC 50 and LC 50 An EC 50 is a statistical estimate of the concentration of a toxicant in the ambient medium necessary to produce a par­ ticular effect in 50% of a very large population under speci­ fied conditions. An LC 50 is a special case of the EC 50 in which the effect recorded is death. EC 50 and LC 50 values are determined similarly to ED and LD 50 values, that is, probits representing the percentages of animals showing a response in a series of tests are related to the logarithms of the concentrations that produce the responses. EC 1 and EC 99 values may be determined and confidence limits of the vari­ ous estimates may be calculated. ET 50 and LT 50 values may be calculated on the basis of concentration instead of dosage.

1.3.5  Measurement of Graded Responses What has been said so far about quantitation of dosageresponse relationships was concerned with all-or-none effects or effects that can be treated in this way. However, many

A simple but profound change in pharmacology began in the 1940s when increased emphasis was placed on the impor­ tance of tissue levels of drugs. It had long been known that chemicals act at the cellular level, and the change in empha­ sis was conditioned largely by the rapidly increasing ability to carry out the necessary measurements which made it pos­ sible to relate the effectiveness of many drugs to their mini­ mal plasma levels or, to be more exact, to minimal plasma levels of free compound. This critical index of the concen­ tration available to cells made it possible to devise rational, nontoxic uses of several compounds that previously had been too slowly effective or too toxic to be practical. The factors that can be involved in determining the plasma levels of free compounds have been reviewed by Brodie (1967) and by the National Academy of Sciences (1969). Some factors other than dosage that control the concen­ tration of foreign compounds available at the tissue level are discussed later. Ultimately, the concentration of a drug or another (bio)chemical moiety at the site of action determines its toxicity, even for hit-and-run-type poisons whose continu­ ous presence is not required. Therefore, the perfect EC 50 or LC 50 value would be to determine such a number directly in the target organ or tissue. However, that is often very difficult or impossible, because it requires either tis­ sue biopsies or killing of the animals. Physiologicallybased pharmacokinetic modeling eventually could become a very useful tool because it can predict tissue concentra­ tions, but its accuracy is as yet unsatisfactory. In practice, toxicologists use surrogate measures for the concentration at the site of action. The most widely used, but not the best surrogate is the dose. The best surrogate is the plasma con­ centration of a chemical because the free fraction of the agent in plasma is in equilibrium with the free fraction in the target organ or tissue. It is often very difficult or even incompatible with the experimental design to obtain blood repeatedly from experimental animals. Therefore, the dose will remain a useful surrogate if properly qualified. The dose will be a poor surrogate for the tissue concentration of cadmium in testes after oral administration, because of very limited absorption from the gastrointestinal-tract, but it still will be proportional to it. Dermal absorption of some

34

chemicals is almost nil or very limited when high doses are applied to the skin. In such instances the dose will be a poor surrogate if at all for tissue concentrations. In inhala­ tion studies the dose is again a good surrogate for the con­ centration at the site of action without regard for whether the lungs themselves or distant organs or tissues are the target(s) of toxicity for both types of compounds whose absorption is ventilation- or blood-flow-limited. Often the concentration of a volatile compound in the inhaled air is used as a surrogate. It is easy to convert an inhalation con­ centration to a dose if the body weight is known because the physiological parameters of respiration are well estab­ lished in humans as well as in laboratory animals. There are some nearly perfect surrogates for target tissue levels such as the concentration of a water-soluble chemical in an aquarium or in an in vitro experiment.

1.3.7  Statistical Considerations There are several good reference books on statistics appli­ cable to problems in toxicology. These include volumes by Pearson and Hartley (1976), Mainland (1963), Steel and Torrie (1980), and Snedecor (1967). Useful statistical tables may be found in books by Beyer (1968) and Fisher and Yates (1963). A book designed specifically for toxi­ cologists is that of Gad and Weil (1986) and a more recent one by Gad (2006). This section is not intended as a substitute for the ref­ erences just cited, and such books must be consulted by anyone interested in the mathematical details. This section does discuss some broad guidelines regarding (a) how many subjects are required for ordinary tests, (b) randomization of subjects, (c) selection of dosage levels, and (d) species con­ siderations associated with the effects of small dosages.

1.3.7.1  Number of Subjects As discussed in Section 1.3.1.1 and illustrated graphically in Fig. 1.4, the accuracy of statistical mea surement can be increased by running more tests under the same conditions. Number of Independent Units  Table 1.4 shows the differ­ ence that must appear between two equal groups to be signifi­ cant at a level of P  0.05, using groups of 50, 40, 30, 20, and 10 subjects, respectively. The simplest solution is that in which the effect under study does not occur in the controls. Inspection of the table shows that, when no controls are affected, there still must be 5 reactors (50%) among a group of 10 experimental subjects to achieve assurance that the dif­ ference between the two groups has not occurred by chance. If the groups are larger a smaller proportion of reactors is required. Thus, with groups of 50 subjects each, only 6 reactors (12%) are needed in the experimental group to indicate a sta­ tistical significance of difference when no controls react. Put

Hayes’ Handbook of Pesticide Toxicology

another way, groups of 50 subjects each will be required to demonstrate dependably an effect that occurs in 12% of a very large population even when there are no reactors among the controls. Larger groups are required if events that occur at a lower frequency are to be demonstrated. The second and succeeding lines of Table 1.4 are con­ cerned with the situation in which there are reactors among the control group. If one group has a certain percentage of reactors, the other group must have a specified larger per­ centage for the difference to be significant. Thus, at least 50 animals per group would be necessary in order to give reasonable assurance that 26% incidence in one group and 10% incidence in the other group are in reality different. Table 1.4 is intended to illustrate the principles just out­ lined. A more complete table suitable for guiding experi­ mental work is provided in Mainland (1963). Needless to say, even a mild clinical effect of a com­ pound would be intolerable if it occurred in 1% of the gen­ eral population who encountered traces of the material. If the effects were at all serious, an incidence of 1% among workers would be intolerable also. The solution of the problem from the standpoint of ordinary testing is to keep the limitations of precision in mind and to design experi­ ments and select dosages in such a way that there will be one or more groups in which the parameter of interest approaches an incidence of 100%, while the incidence in the control is held very low. Interpretation of the results must be based not only on statistical consideration, but also on judgment regarding severity and reversibility of the effect under discussion, and the relevance of the test as a whole to the human situation. Identity of Sampling Unit  As critically reviewed by Weil (1970), it is an error to count individual subjects as statisti­ cal experimental units when these subjects are not randomly selected. For example, in studies of reproduction or terato­ genesis, mothers (or litters) and not the number of offspring are the proper basis for statistical analysis. It is misleading to report the number of litters showing any malformation or, more precisely, the proportion of malformed young per lit­ ter. The reason, of course, is that the fate of any particular offspring is conditioned by the physiology of its mother and by the dosage she received. Counting young rather than litters counts the same thing over and over. This tends to exaggerate the statistical significance of the results and may lead to the conclusion that observed differences are significant when they easily might occur by chance. The same precaution must be observed in connection with studies of carcinogenicity started with newborn or infant animals. Other types of unjustified grouping for statistical analy­ sis are the combination of dosage groups, sexes, or strains

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

35

Table 1.4  Difference between Two Groups Necessary for Significancea 50 animals per group

40 animals per group

30 animals per group

20 animals per group

10 animals per group

Least affected

Most affected

Least affected

Most affected

Least affected

Most affected

Least affected

Most affected

Least affected

Most affected

0

12

0

15

0

20

0

30

0

50

4

16

5

20

3 1/3

26 2/3

5

35

10

70

10

26

10

27 ½

10

36 2/3

10

45

20

80

20

38

20

42 ½

20

46 2/3

20

55

30

90

30

50

30

52 ½

30

56 2/3

30

65

40

100

40

60

40

62 ½

40

66 2/3

40

75

50

100

50

70

50

72 ½

50

76 2/3

50

85

60

—

60

80

60

82 ½

60

86 2/3

60

90

70

—

70

88

70

90

70

96 1/3

70

100

80

—

80

94

80

95

80

100

80

—

90

—

90

—

90

—

90

—

90

—

a

From National Academy of Sciences (1960), by permission of the Academy. Differences are given as percentage incidence. P  0.05.

without testing the data statistically and finding that a par­ ticular combination is justified. A number of pitfalls in the applications of statistics have been discussed by a Task Force of Past Presidents (1982).

1.3.7.2  Randomization of Subjects There is a possibility that error will be introduced into exper­ iments through nonrandom selection of subjects, whether animals or people. For example, in selecting animals from a holding cage, there is a chance that the quieter ones will be taken first and the livelier ones caught later. If they are placed in groups as they are caught, there will be less ten­ dency for successive groups to be more active, and the degree of activity may represent a basic physiological differ­ ence. The remedy is to give the animals temporary numbers as they are caught and then assign them to groups accord­ ing to a table of random numbers. Such tables frequently are included in books of statistical tables. After the animals have been divided into groups, each one may be given a perma­ nent identification number in serial order. What has just been said about randomization concerns populations that either are considered homogeneous or exhibit variation impractical to control. Of course recog­ nized variation may be made the object of controlled exper­ imentation. For example, if differences involving sex are to be studied, the males and females must be segregated, after which subjects of each sex may be assigned randomly to

appropriate groups. Because populations may lack homo­ geneity in many obvious ways, it is often desirable to limit a series of tests to animals of a preselected age or weight. In working with a limited supply of subjects, it maybe better to ignore strict randomness in order to distribute some obvious variables among the different experimental groups. For example, if 30 men of widely different ages are to be placed in three dosage groups, it may be desirable to place their names on cards, arrange the cards in order by year and date of birth, and then deal the cards into three groups in the order 1, 2, 3, 3, 2, 1, 1, 3, 2, and so on. Although such a distribution is not random, it will eliminate bias. After a test has been run in this way, the data may be examined to see whether recognized variables affected the result. The results for each group may be plotted by age, or a coefficient of correlation may be calculated to see whether the outcome was significantly influenced by this variable. The results for one race may be compared to those for another, and other variables may be considered in turn.

1.3.7.3  Selection of Dosage Levels Because the effects of chemicals expressed as probits usu­ ally form a straight line when plotted against the logarithm of dosage, it is best to choose a series of dosage levels that form a geometric progression. A factor of 2.0 (log inter­ val of 0.3) is often used. More detailed information will be obtained if a smaller factor such as 1.26 (log interval of 0.1)

36

is used, especially in the region of the ED 50. Conversely, less (but sometimes sufficient) information can be obtained by testing dosage levels separated by a factor of 4 or more. Selection of the general range of dosages to be studied is simply a matter of judgment supplemented by cautious trial. A number of methods for efficient use of small numbers of animals for determining ED 50 and LD 50 values are refer­ enced in Section 1.3.1.1. Use of the logtime-logdosage curve for predicting the proper dosage range for tests involving repeated doses is discussed in Section 1.3.3.2. These guidelines refer to quantitative studies of chemi­ cal effects that are agreed to exist, such as the lethality of any compound if given to any species at a sufficiently high dosage. However, for scientists who do not fully accept the dosage-response principle, a special problem in the choice of experimental dosage levels is presented by any study of a property that is thought to be possessed by some com­ pounds but totally lacking in others. For example, use of the highest tolerated dosage is common in testing for car­ cinogenicity (Section 1.4.4). This is justified by those who practice it as providing the greatest statistical possibility of revealing a positive result. It is criticized by some others as providing an unrealistic result because it is based on dos­ age levels that people are unlikely to encounter and even experimental animals are unlikely to be able to detoxify and eliminate as effectively as they would the levels that occur in the environment. It seems likely that no solution to this dilemma will be reached except by relating the results of animal studies to epidemiological investigations. The point is that no comparable problem in the choice of experimental dosage levels exists in connection with a study of a property such as lethality that is common to all compounds if administered at sufficient dosage. Thus a demonstration, whether in animals or humans, that table salt can he fatal does not pose any difficulty in evaluating its proper use. On the contrary, a demonstration that even a high dosage of a compound is carcinogenic in any species compromises the evaluation of it by those who do not fully accept the dosage-response principle. Although it seems likely that there are compounds that have and others that lack specific properties and although these differences may involve basic toxicological differences, such as the appar­ ent lack of a third segment of the logtime-logdosage curve by carcinogens (Section 1.3.3.2), the main problem in eval­ uating compounds that do not have specific objectionable properties is one of established patterns of human thought, not of science. If this were not true, every potentially lethal compound would have to be banned, which means every and all drugs and other chemicals. Statistical considerations did and still do play a predomi­ nant role in designing toxicological experiments. Indeed in the absence of a theory, which Hayes (1991) lamented in the discussion of the next section, proper statistical analysis is the only way to distinguish between chance occurrence and a cause-effect relationship usually characterized as 95% or

Hayes’ Handbook of Pesticide Toxicology

higher confidence that two or more frequency distributions are truly different. Often, large biological variability is being made responsible for the lack of finding significant effects at some low dose in a defined population. The possibility that the standard toxicological protocols might be the primary reason for large variability for intra- and inter-laboratory results was not entertained, at least not from a theoretical point of view. Most ED or LD studies conducted in the past supposedly measured toxicity, when in fact they measured different combinations of intoxication and recovery. Because two or more variables were changing at the same time, the biological variability was greatly magnified. A third tradi­ tionally uncontrolled variable was due to the lumping of dead animals that died of different causes (mechanisms). Vehicle, formulation, volume, etc. are additional variables, which for the most part were much better controlled than these two much more important factors identified previously. The most reproducible studies were the ones conducted at toxicokinetic and/or toxicodynamic steady state, which “nature” some­ times happened to provide in the form of long kinetic and/or dynamic half-lives. For example, a compound with a kinetic or dynamic half-life of 1 h and a time to effect of 10 min will provide nearly pure toxicity data as does a chemical with a kinetic or dynamic half-life of 1 year and a time to effect of 60 days, because in both instances recovery will be relatively insignificant during the period of observation. However, as the ratio of kinetic or dynamic half-life to observation period becomes less and less favorable the contribution of recov­ ery becomes greater and greater. It must be understood that just because an animal happens to die does not mean that the organism did not try to recover from the toxic insult. All definitive toxicological experiments have to be conducted under kinetic or dynamic steady state (ideal conditions) to determine the respective toxicological constants. Having done that makes the dose selection highly accurate accord­ ing to c  t  k  W. Plotting c  t vs. W yields a straight line with the slope 1/k (Rozman, 2000a). Doing toxicologi­ cal experiments “right” would have several advantages: They would require fewer animals, because of reduced variability. The experiments would become more reproducible; in fact, if conditions were kept ideal they would become entirely repro­ ducible. The erroneous conclusion that the relative potency (structure activity by the same mechanism) of a chemical is dose-dependent (Fig. 1.5) would also be eliminated.

1.3.7.4  Effects of Small Dosages Safety evaluation is much concerned with the effects of dos­ ages just below and just above the threshold of observable effect, that is, with the no-significant-effect level and lowest effect level in practical experiments. After a consideration of these practical matters, the following paragraphs will pres­ ent a discussion of what is known about the existence of thresholds and the beneficial effects of small dosages. These theoretical matters have clear implications for the probable

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

outcome of the quantitative study of the effects of dosages so small that extremely large groups of subjects are required for their meaningful investigation. Because of its diffi­ culty and expense, this kind of quantitative study has rarely been carried out (Section 1.3.1.3), but further study would have tremendous theoretical and practical importance. The strictly toxicological study ought to be coordinated with biochemical investigations that could offer a reason for the statistical findings, no matter what they may be. The No-Significant-Effect Level  The concept of “no effect” is one of the most commonly employed in safety evaluation. The no-observed-effect level is the maximal or near maximal dosage level at which no difference from untreated or vehi­ cle-treated controls is detected. It is not a level so far below any effective one that it is insignificant. Although the term “no effect” often is applied in connection with a dietary con­ centration, it is always in the context of the dosage admin­ istered. Although the term could be employed in connection with one or a few doses, it is employed mainly in connection with long-term tests. Use of the expression implies an orga­ nized study and careful observation, not the casual result of a test with some other purpose. Any study designed to reveal a dosage producing no effect should contain, in addition to a control group receiving no toxicant, one or more groups of subjects given larger dosages fully expected to produce measurable effects of the compound. Because such tests are almost always prolonged, particular attention must be given to the number of subjects assigned to each group so that the number available at the end of the study will be adequate to reach a statistically significant result (Section 1.3.7.1). There are several reasons for placing the expression “no effect” in quotation marks. First, many studies reveal effects that are obvious, reproducible, and highly significant from a statistical standpoint, but of questionable biological sig­ nificance. Depending on circumstances, an example might be partial inhibition of an easily regenerated enzyme. There is no substitute for judgment in toxicology. Second, as dis­ cussed in one of the following sections, the effects of small dosages of toxic compounds may be beneficial and thus qualitatively different from the effects of larger dosages. Third, failure to detect any effect by an elaborate scheme of testing does not exclude the possibility that an effect would have been detected if some other scheme had been used. Rapid progress has been made in chemical detection of toxicants or their metabolites. Analytical chemists already have achieved such skill that they easily measure biologi­ cally unimportant traces of several pesticides. Biochemists and physiologists are not far behind. For these reasons, it is generally recognized that what is needed is a no-significant-effect or no-observed-adverseeffect level (NOAEL), that is, a level that causes no detect­ able injurious effect. There is no complete substitute for long-term tests, but increasing attention must be given to evaluating the biological significance of observed effects

37

that involve no demonstrable injury to health. A toxicolo­ gist may do as much harm by unnecessarily condemning a compound as by failing to detect and prevent a real toxic hazard. Prophetic words indeed, albeit no attention was paid to them. In the contrary, Dr. Hayes’ concern and fear has become today common practice among many toxicologists. The Lowest Effect Level as a Practical Toxicological Measurement  As already mentioned, it usually is imprac­ tical to use such a large number of subjects per group that the possibility of a rare (1%) but highly undesirable effect, such as neurotoxicity or carcinogenesis, can be excluded in the lower dosage range. This problem is not as important as it may appear at first, because the frequency of an effect can be increased by increasing the dosage (just one reason for using injurious dosages in long-term stud­ ies). Thus, the existence of a highly undesirable effect of a particular compound can be taken into account in the selection of a safety factor. A more serious objection from a purely scientific standpoint is the imprecision of finding the highest no-significant-effect level. From a statistical point of view, it would be preferable to employ for safety evaluation the lowest effect level, that is, the smallest one at which a meaningful, injurious effect or even a relevant, harm­ less effect is detected. This implies that (a) the no-significanteffect level is determined in order to put a limit on what is meant by the smallest effect level, but (b) the more objec­ tive, positive finding is used for safety evaluation, and (c) the nature of any injurious effect observed will be taken into account in choosing a safety factor. Threshold Levels as Biological Facts  The practical dif­ ficulty in establishing a “no-effect” level for a particular compound using a manageable number of experimental ani­ mals and the more complex problem of extrapolating a safe level for humans must not be permitted to obscure the fact that thresholds do exist. The toxicologist faced with a single limited experiment would be wise to recall the impossibility of proving a negative. The logician faced with a complete lack of supporting evidence would be wise not to press pure logic too far and conclude that no threshold exists so that even a single molecule may represent a hazard. As pointed out by Friedman (1973), the question of the existence of a threshold is a problem of biology, not of mathematics or of probability. In not a single instance has the absence of a threshold been demonstrated. On the contrary, concentra­ tions are known at which compounds with the highest bio­ logical activity are inactive. For example, as described by Friedman (1973), limita­ tion of vitamin A intake to 10% of the minimal required dosage leads to severe deficiency disease, yet this consti­ tutes a dosage of 3.6  1015 molecules per kilogram of body weight or a concentration of 6  109 M in the body. In a similar way, ineffective levels of vitamin D and vitamin B12 (for which the daily requirements are only 10 g/day and

38

1 g/day, respectively) are 4  1011 M and 1  1012 M. By conservative estimates, post-menopausal women and adult men have 1.5  l012 molecules of estradiol per kilo­ gram of body weight or the equivalent of 2. 6  1012 M. For 2,3,7,8-tetrachlorodibenzodioxin (TCDD), reported to have an LD 50 value of 0.0006 mg/kg in guinea pigs, a harmless dose would still produce a concentration of 1.9  1010 M. Botulinum toxin, for which activity described in terms of molecules has long been common, would pro­ duce absolutely no effect in mice at a dosage of 4.2  107 molecules per kilogram or a concentration in the mouse of 7  1017 M, assuming a molecular weight of 900,000. Values of the same orders of magnitude apply to some carcinogens. An ineffective amount of aflatoxin in the rat consists of a dosage of 9.6  1012 molecules per kilogram and a concentration of 1.6  1011 M. However, many strong carcinogens are less potent. For 1, 2, 5, 6-­dibenzanthracene, methylcholanthrene, and 3,4-benzpyrene administered by different routes, the ineffective concentration in the body ranges from 1  108 to 1  101 M. The limiting level is even higher for compounds that are not inherently very active. For example, an ineffective amount of Aramite as a carcinogen is a concentration of 3  101 M. Hutchinson (1964) and later Dinman (1972) sug­ gested that 104 molecules per cell is the limiting concen­ tration for biological activity, whether pharmacological or injurious. As pointed out by Friedman (1973), there are about 6  1013 cells in a 70-kg human body, from which it follows that the suggested limiting level for activity is 8.6  1015 molecules per kilogram or about 1  108 M. The demonstrated no effect levels for vitamin D, vitamin B 12 and estradiol (1011–1012 M) are so low that the cor­ responding thresholds may be somewhat lower than the limiting level of 1  108 just discussed. The same reason­ ing applies to the thresholds of toxic action of TCDD and botulinum toxin and the threshold of carcinogenic action of aflatoxin. Further evidence that the limiting concentra­ tion for the inactivity of a few highly active compounds is less than 108 M is the report that certain pheromones have thresholds on the order of 1  1012 M. These values do not prove that the limiting concentration is even less than 108 M in susceptible cells, because some compounds, such as botulinum toxin, have extremely high affinity for the cells on which they act, and their distribution in the body at critical dosage may be very uneven. However, exactly what concentration is limiting is far less important than the fact that thresholds do exist even for the most active compounds. No one doubts that the existence of deficiency conditions proves that minimal or threshold concentrations of vitamins, hormones, and other beneficial compounds are required for proper action. There is no chemical or other scientific reason to suppose that there is an inherent, fundamental difference in the dosageresponse relationship of injurious compounds.

Hayes’ Handbook of Pesticide Toxicology

Predicting the effects of small doses has been one of the core problems of toxicology (Rozman and Doull, 1998). Knowledgeable toxicologists were always aware of the existence of biological threshold doses, which would not cause any response in a given biological system. However, because of a statistical rather than a theoretical or biologi­ cal view of the dose-response and because of a lack of considerations of time as a quantifiable variable of toxic­ ity, a definition of a threshold in toxicology remained elusive. As long as one looks at toxicity only in terms of dose-responses it is logical to arrive at NO(A)ELs and LO(A)ELs as starting points for safety and risk assess­ ments. The problem is that a NOEL is fuzzy if only a few doses have been used, often one or more orders of magni­ tude apart. The LOEL is less fuzzy because the magnitude of the effect provides an estimate of how far away a NOEL might be for a given population size. Here toxicology has encountered a thus far insurmountable difference of opinion among its practitioners with no resolution in sight. There are those who believe that some dose-responses are so shal­ low at the lower end that the terminal slope is essentially linear, which corresponds to a very large standard deviation in terms of frequency distribution of a normally distributed effect. Others point out that the concept of the maximum tolerated dose (MTD) prevents the experimental exploration of a full dose-response and that most of the currently avail­ able long-term studies represent truncated dose responses limited to the low dose end of the dose-response. A lack of theoretical considerations is at the heart of this difference in opinion. Most chronic studies (acute experiments were argued in Section 1.2.2.1) entail intermittent administration of chemicals with different periods of recovery between exposure episodes. If the kinetic-dynamic half-life (as is the case for most chemicals) of a compound is very short then there will be slow if any accumulation of injury because of significant or complete intermittent recovery. Thus, most of the chronic toxicological experiments of the past mea­ sured various combinations of intoxication and recovery. Dependent on the ratio of the two (and on additional “hid­ den” variables) the frequency distribution and hence the dose response becomes flatter and flatter because individu­ als almost never belong to the same normal distribution in terms of both injury and recovery. For example, the gene responsible for producing acetylcholinesterase (AChE) in the respiratory center is different from the genes produc­ ing carboxyesterases released into blood. Individuals could have a high synthesis rate of AChE in the pons and medulla but low production rate of carboxyesterases or vice versa in any quantitative ratio. Susceptibility of an organism to organophosphate toxicity depends on the carboxyesterase pool (detoxification) but recovery from the intoxication is determined by the rate of AChE production in the critical brain region. The conclusion from these considerations is that to measure true (pure) toxicity one must conduct an experiment under kinetic-dynamic steady state, which often

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

is possible only by continuous exposure until the occur­ rence of a well-defined effect. Although it might be often impractical to live up to this ideal condition, experimental­ ists must be aware of it to avoid claiming exceptions from the theory of toxicology when, in fact, they are not control­ ling some variables. Experiments conducted under ideal conditions controlling all variables but the dose and one timescale (that of intoxication) will always yield c  t  k under isoeffective conditions or c  t  k  W under iso­ dosic or isotemporal conditions. It is from this relationship that a true LO(A)EL can be determined by substituting for time (t) the maximum life span of a given species. In its immediate vicinity is the NO(A)EL in terms of threshold dose which will cause no effect whatsoever in a lifetime, for a given population, cthreshold 

k W tlifespan

cthreshold 

(isodosic and isotemporal) k tlifespan

(isoeffective)

Beneficial Effects of Small Dosages  Schulz (1888) may have been the first to observe stimulation from very low concentrations of poisons. He investigated the effect of mercuric chloride,iodine, bromine, arsenous acid, chromic acid, formic acid, and salicylic acid on yeast and concluded that, when sufficiently diluted, all of them can increase the vitality of yeast over a longer or shorter period of time. Only a few years later the bacteriologist Hueppe (1896) stated the rule that has come to bear his name. “Every sub­ stance that kills and destroys protoplasm in certain con­ centrations inhibits development in lower concentrations, but acts as a stimulus and increases the potential of life at even lower concentrations beyond a point of neutrality.” In stating this principle, Hueppe mentioned certain apparent exceptions. He also acknowledged the independent discov­ ery of the rule by Arndt and Schulz. A special universally accepted instance of the beneficial effects of small dosages involves the reactions of plants to what are now called essential trace elements. A plant can­ not live if even a single one of these metals or metalloids is absent from the medium in which the plant is grown, but an excess of any one of them is injurious. This is the law of optimal nutritive concentration. Recognition of it must be credited to Gabriel Bertrand even though he may never have stated it concisely in a publication. The relationship of beneficial and harmful effects of trace elements was implied in the discussion that followed his presentation on “complementary nutrients” at the Fifth International Congress of Applied Chemistry held in 1903 in Berlin (Bertrand, 1903). Personal communications from his son, Dr. Didier Bertrand, and one of his students, Dr. Rene Truhaut, established that Gabriel Bertrand presented the

39

law of optimal nutritive concentration in his course at the Sorbonne from 1908 through 1930. Later, M. D. Bertrand (1962a, 1962b, 1969) generalized the law and expressed it in a mathematical form. As reviewed by Townsend and Luckey (1960) and in a very different way by Smyth (1967), evidence has contin­ ued to accumulate that small dosages of many compounds are not injurious, but beneficial. Townsend and Luckey tabulated many examples from the pharmacological litera­ ture. Smyth offered several original examples of benefits from small doses of toxicants. The benefit may be substan­ tial and include increased rate of growth, greater fertility, and prolonged life span. The phenomenon, or variants of it, has received different names. Thus the noun “hormesis” and the adjective “hormetic” were proposed for the stimulatory action of a subinhibitory amount of a toxicant (Southam and Ehrlich, 1943). The term “hormoligosis” (from the Greek hormao, rouse or set in motion, and oligos, small) was pro­ posed (Luckey, 1956) to indicate the more general process by which a small amount of anything, regardless of its tox­ icity, produces stimulation. The same author used the term “hormoligant” to indicate something that stimulates when given in a small amount. The term “sufficient challenge” introduced by Smyth (1967) refers to the entire range of phenomena and emphasizes the need of the organism for some measure of stress, whether it be a small amount of poison, a small amount of radiation, or early, immunizing infection. In fact, he points out that he took the term from Toynbee’s concept of “sufficient but not overpowering chal­ lenge” in connection with human history. There is a tendency to take for granted the beneficial effects of small amounts of certain classes of compounds, which we call drugs, nutrients, or growth promoters, and to ignore completely the beneficial effects of small amounts of other materials, some of which we call poisons. The dif­ ference depends largely on our supposed understanding of their actions. Since antiquity, the use of therapeutic drugs has seemed reasonable to people. Thus we “understand” the benefits from this one class of materials that are clearly toxic at higher doses. A nutritional mechanism for the stimulation produced by low concentrations of certain toxic substances offers another basis for understanding. A number of minerals, vita­ mins, amino acids, and fatty acids are known to be essential to animals. The fact that excessive intake of some of them, notably several of the metals and vitamins A and D, has led to cases of human poisoning, has not detracted from acceptance of their benefit. The discovery of the induction of processing enzymes, especially the mixed function oxidases of the liver, has added a third means of understanding the benefit of small amounts of some drugs, pesticides, and other chemicals. Finally, the effectiveness of growth-promoting feed additives is understood in a somewhat different sense. There is little or no reason to think that the effectiveness of arsenilic acid and various antibiotic feed additives depends on any nutritive

40

value. Their mode of action when given in the usual homeo­ pathic dosages is suspected only to be related to action on the microflora, but their ability to make chickens, pigs, and calves grow faster is inescapable. In this case understanding is mainly in terms of commercial success. There is a tendency to ignore the beneficial effects of small doses of toxic compounds unless they are understood in terms of therapeutic action, nutritional requirement, growth promotion, or perhaps enzyme induction. In the lat­ ter case, there is some ambivalence and tendency to view adaptive change as evidence of injury. A few scientists have the courage to see in the induction of enzymes an evidence of adaptation at the molecular level. Toxicologists should combat all bias. The existence of a phenomenon does not depend on our understanding of it. Statistically established evidence of benefit from small dosages ought to be viewed just as objectively as statistically established evidence of injury from larger dosages. This statement is not meant to underestimate the importance of increasing our basic understanding: it is a plea to explore widely and to accept facts even when they appear contradictory. Most compounds have two or more modes of action which reach expression at different, though perhaps over­ lapping dosage levels. Possession of more than one mode of action certainly opens the possibility that low dosages of a compound will be beneficial rather than merely harm­ less. However, both benefit and harm may be associated but are not necessarily associated with a single mode of action. Most, but not all, of the side effects of drugs are excessive expressions of their therapeutic actions and the result of overdosage. It is a general principle that excessive dosages of ben­ eficial compounds are always toxic. It may be that the con­ verse is also true, for the possibility cannot be excluded that sufficiently small dosages of toxic compounds are always beneficial in some living system: each apparent exception may be merely the result of failure to test a par­ ticular material under appropriate conditions. Hayes (1991) was fully aware of the widespread pres­ ence of beneficial effects of small doses of chemicals such as drugs, essential nutrients, and other hormetic agents. Calabrese and Baldwin (2001a) evaluated the literature and found hundreds of cases of clear (dose-dependent) hor­ metic effects of chemicals that otherwise were only looked upon as toxicants. In contrast to Hayes (1991), who took an integrated view of the beneficial effects of chemicals (with which we agree), Calabrese and Baldwin (2001a) assumed a viewpoint of conceptual fragmentation by excluding drugs and nutrients from the group of chemicals having hormetic effects. The  curve suggested by Calabrese and Baldwin (2001b) to conceptualize hormesis is also problematic. It is implausible from the biological point of view that a chemi­ cal can cause both an increase and a decrease in an effect by the same mechanism (Rozman and Doull, 1999). It is more reasonable to extend Hayes’ view to all chemicals having

Hayes’ Handbook of Pesticide Toxicology

the potential of exerting beneficial (hormetic) effects in small doses, although often the benefit may be immeasur­ ably small. Increasing doses will neutralize the benefit by a different mechanism and eventually lead to toxicity by the same or a still different mode of action. Thus the  curve is an attempt to combine two or three parallel or nonparallel dose responses into a single curve. It has no foundation in the principles of toxicology. Homeopathic claims of small doses of naturally occurring or man-made chemicals are equally incompatible with the principles of toxicology unless supported by clear dose- and time-response relationships. In fact, the principles of toxicology (large doses) and pharmacology (small doses) are highly compatible with Toynbee’s view (see Smyth, 1967, and Hayes above) of successful and unsuccessful societies in historical context, which means in the dimension of time. A “sufficient, but not overpowering challenge” which increases the fitness of a society to respond to larger challenges is similar to small doses, which enhance the fitness (adaptation) of an individ­ ual to respond to subsequent higher (toxic) doses. However, an overpowering challenge very much like a very large dose results in the demise of both a society and an individual. Logprobit Model and Quantitative Study of the Effects of Small Dosages  It is implied by the logprobit curve for dosage-response that an effect occurring in a low pro­ portion of a population, for example, an average of 1 in 10,000, is merely the result of a smaller dosage than what would produce the same effect in a high proportion of the same population. The number of subjects necessary to measure the effect of a truly small dosage is very great unless this effect is qualitatively different from that pro­ duced by a large dosage. If a small dosage produces a beneficial effect com­ pared with the control, it may be possible to establish this fact with no more experimental subjects than are required to show that a larger dosage is harmful, and the presence of a positive benefit may exclude the possibility of a hypo­ thetical injury from the same dosage. This principle is illus­ trated by facts regarding selenium, a trace element essential to life. Using groups of only eight animals each, G. Siami (personal communication, Siami, 1971) showed clearly that rats fed a dietary supplement of 1 ppm selenium grew better than rats fed the same regular commercial rat feed which contained only 0.15 ppm of the element. Rats receiving a supplement of 2 ppm also grew better than the controls but not as well as those receiving the 1-ppm supplement. Rats receiving 3-ppm supplement showed definite toxicity, including liver cirrhosis, and those receiving 5 ppm of sele­ nium died in only 5 weeks. Thus, the presence of a positive benefit at 2 ppm excludes, within the limits of the experi­ ment, the possibility of injury from this exposure even though 3 ppm was distinctly injurious. If the effects of different dosages are qualitatively iden­ tical and if it requires groups of 10 subjects to identify

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

41

9

99.9900 99.9000

8

99.0000

7

90.0000

6 5

1/2

50.0000

1/10

10.0000

1/100

1.0000

1/1,000

0.1000

2

1/10,000 1/100,000

0.0100 0.0010

1

1/1,000,000

0.0001

4

Probits

Incidence (fractional)

Incidence (%)

99.9990

3

0.001

0.01

0.1

1.0

Dosage (mg/mouse)

•

Figure 1.13  Incidence of tumors in mice following a single subcutaneous dose of 20-methylcholanthrene: ( ) observed values; () values adjusted from 0 or 100% by the method of Litchfield and Wilcoxon (1949). Hypothetical deviations from the straight line logprobit relationship are ahown by a dotted and a dashed line, as discussed in the text. Data from Bryan and Shimkin (1943).

with acceptable precision the dosage that, on the average, causes an event once in every two chances (ED 50), then it would require groups of 500 subjects to find with the same precision the dosage that, on the average, causes an event once in 100 chances (ED 1). For an event affecting 1 in 10,000 (ED 0.01), groups of 50,000 subjects each would be required. Such large groups of subjects are required to measure the effects of small dosages directly that this mea­ surement is entirely impractical in connection with routine toxicological testing. In fact, such studies have been done only rarely. Reasons for not doing statistical studies of small dos­ ages include (a) their great expense, (b) the technical dif­ ficulty of preventing or even recognizing effects caused by uncontrolled variables, and (c) the fact that tests employing a reasonable number of animals per group give results of practical value entirely suitable for determining safe levels of human exposure. The safety factor [whether in the form of a fraction of the no-significant-effect level or the lowest effect level or in the form of standard deviations] removes the acceptable dosage one or two orders of magnitude from the lowest dosage tested experimentally and, therefore, far from the area of danger. The main reason for doing thorough, direct, statistical studies of the effects of small dosages is that we lack that knowledge. For practical purposes, we can compensate for our ignorance by using safety factors that exceed the degree of uncertainty involved. However, there might be practical as well as theoretical value in exploring with precision the

effects of small dosages. Sound information would reveal the limits of variability and thus indicate more accurately what safety factors really are needed-a practical result, greatly to be desired. The linear logprobit dosage-response curve was devel­ oped on the basis of observed facts but also involves the theoretical assumption that response to dosage follows a lognormal distribution. The logprobit model fits the observed facts and greatly facilitates their orderly study and presentation (Section 1.3.1). However, there is no way to be sure that the model fits the facts in an area where almost no measurements have been made. Hypothetically, the logprobit curve could deviate from a straight line either upward or downward in the low dos­ age region. Any true deviation would indicate that the distribution for a complete range of dosage levels is not lognormal under the conditions tested. A deviation in the low dosage level would not indicate that the distribution for dosages near the ED 50 level is not essentially lognor­ mal, for the conditions relative to this level and to low dos­ ages may be qualitatively different. The direction of deviation would depend on the nature of the physiological factor responsible for nonnormal distribu­ tion. If there is no threshold for the effect under study so that no dosage, no matter how small, is totally without the effect measured, then the line must deviate upward and approach a horizontal direction in such a way that it will pass through absolute zero located an infinite number of logarithmic cycles to the left, There is no example to illustrate this condition,

42

but it must be considered from the standpoint of logic. If a statistically valid example were found, it would indicate either an unsuspected variable in the experiment or the exis­ tence of a yet undiscovered principle of toxicological action. If the threshold for the effect under study does not lie in the lognormal distribution but at a dosage higher than that predicted by this distribution, then the logprobit line must deviate downward and approach a vertical direction. There are literally thousands of possible examples of this situa­ tion. As often as not, the lower part of a logprobit curve is made up of points adjusted from zero values. The adjust­ ment is made in the faith that a lognormal distribution is involved, even though some of the observed zero values are based on large enough groups of subjects to be statistically likely to give values higher than zero. Of course a number of examples will have to be tested at low dosage levels and with very large groups of subjects before a conclusion about the existence and eventually the frequency of nonlognormal distribution may be reached. If this kind of nonlognormal distribution were demonstrated, the statistical model would have to be adjusted, but no new principle of toxicological action would be indicated. The mechanism would have to be learned in each instance, but possibilities are known. For example, the downward flexure of the curve might corre­ spond to the transition between low dosages metabolized easily by one pathway and higher dosages that overload the normal pathway and involve other pathways also. It appears likely that, if the kind of deviation from log­ normal distribution under discussion exists, it is related more often to the ability of an enzyme to cope success­ fully with low dosage levels than it is related to beneficial effects. Unlike the situation with selenium mentioned ear­ lier, beneficial dosages of many essential elements are one or more orders of magnitude smaller than the smallest dos­ age observed to be injurious. Figure 1.13 illustrates the matters that have just been discussed. The figure shows the proportion of different groups of mice that developed tumors at the site of injection of a single subcutaneous dose of 20-methylcholanthrene as reported by Bryan and Shimkin (1943). It may be seen that the observed values in the area of the ED 50 correspond well to the expected lognormal distribution. The adjusted values for low dosage levels also correspond well, for they have been made to do so. However, each of these adjusted values for low dosages is derived from an observed value of zero. In Fig. l.13, the straight line required by the observed values in the area of the ED 50 has been extended at each end as required by a lognormal distribution of all values. However, in the low dosage area, an upward flexure of the curve has been inserted to illustrate a no-threshold relation­ ship, and a downward flexure has been inserted to illustrate the opposite deviation from a lognormal distribution. Similar reasoning could be applied to the upper portion of the curve, but it would be of no real interest from the standpoint of safety evaluation.

Hayes’ Handbook of Pesticide Toxicology

Statistics is a legitimate and useful tool to describe incompletely understood phenomena as discussed by Hayes (1991). However, overuse of statistics as it occurred and still is occurring in toxicology hampers the development of a theory which represents the only true epistemological gain for a discipline. In the mean time, a large number of statistical methods other than the logprobit method have been proposed (Holland and Sielken, 1993). Yet safety and risk predictions have not become more accurate: they just keep appearing more and more sophisticated mathemati­ cally. Models are only as good as their underlying assump­ tions, which in the absence of a theory are just slightly or not at all better than superstition. Let us now look at the Hayes (1991) example (methyl­ cholanthrene and skin cancer) in light of the theory of toxi­ cology. The kinetics of polycyclic aromatic hydrocarbons has not been studied much, which is surprising considering the enormous amount of work that went into studying their carcinogenicity. Limited data (ATSDR, 1995b) indicate that their biological half-life is on the order of hours to a day, depending on the route of administration. However, a single subcutaneous dose rate (dose) of 20-methylcholanthrene caused up to 80% incidence of cancer at the site of injection, which indicates that this compound’s toxicity is dominated by dynamic rather than kinetic-dynamic processes. This is compatible with a long half-life of the postulated DNAadduct in analogy to benzo[a]pyrene. The actual dose range for measurable incidence encompasses almost 2 orders of magnitude indicative of a shallow slope (Fig. 1.13) associ­ ated with large uncertainty as to the shape of the curve in the low dose region. A single dose rate (dose) study in combi­ nation with a very long observation period is far from ideal conditions because the organism is not at steady state with regard to the injury. Rather the animals are continuously repairing the damage while the cancer is developing. Those animals receiving lower doses have more time to repair the damage, which further increases variability, because some animals actually do reduce the DNA damage to an extent, which does not necessitate the development of cancer within their natural life span. The major conclusion to be drawn from these considerations is that increased variability in the low dose region is due to the concurrent measurement of toxicity (cancer) and recovery. The experiment which does not allow for recovery to occur has also been conducted (Horton and Denman, 1955). Continuous exposure of mice to methylcholan­ threne caused cancer in a highly predictable manner with a c  t  126.7  11.1 mg/kg/week, representing a mere 4.2% variability over a very steep dose (total dose) range of 4.0–15.1 mg/kg. For a compound with a short half-life like methylcholanthrene only repeated exposure will result in the animals attaining toxicodynamic steady state after about 4 recovery half-lives after which c  t  k occurs accord­ ing to a rectangular geometry. Mistaking the (daily) dose rate for the dose, when, in fact, the dose is always the sum

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

of (daily) dose rates, has been one of the most damaging assumptions toxicologists ever made, because it makes the c  t  k concept disappear as was the case with methyl­ cholanthrene. It must be recognized that assuming that a daily dose rate is a dose amounts to ignoring time, in that animals receiving lower doses live a lot longer than those given high doses and therefore the daily dose rate artificially and with no theoretical legitimacy flattens the slope of the dose response, leading to additional variability in the low dose region. The fact that no mouse lives beyond 1500 days allows one to conduct a highly accurate biological safety-risk assessment by substituting cthreshold  1500  126.7  11.1 yielding a dose for cthreshold  0.85 mg/kg or a daily dose rate of 0.004 mg/kg/week. The uncertainty about this number arising as a result of the variability of k (4.2%) is very small (Section 1.3.7.4). In other instances, the relationship between applied dose and the time to tumor was unsatisfactory even when plotting dose rather than dose rate (daily dose) vs time to effect as with benzo[a]pyrene (Poel, 1959). As discussed previously in Sections 1.1.3 and 1.1.4. there are two important vari­ ables that need to be controlled or otherwise the c  t  k relationship is lost. One is the exact definition of the effect such as time to onset of a response (first tumor) or time to 100% response (single tumor) because the definition of ideal conditions requires that experiments examining dose-time relationships must he conducted under isoeffective condi­ tions. This was not the case with benzo[a]pyrene, but with methylcholanthrene it appears to have been the case. The other problematic variable in toxicity studies using dermal application is that the applied dose is often a poor surrogate of the dose (concentration at site of action), because after saturation of the flux through a defined skin surface, there is no more proportionality between applied dose and systemic dose. It is quite apparent that neither variable was wellcontrolled in the study with benzo[a]pyrene and, perhaps fortuitously, both were controlled in the experiment with methylcholanthrene. This discussion illustrates that improvements in safetyrisk assessment are not going to arise out of more statistical sophistication, but out of conducting toxicological experi­ ments under ideal conditions (kinetic-dynamic steady state) and then understanding how departures from ideal condi­ tions influence the slope of the dose-time- responses and their variability at low doses and short times. Other Models of Dosage-Response Relationships  Other ways of describing dosage-response relationships include the logprobit and one-hit models. They give results very similar to those for the logprobit method of analysis at the ED 50 level and even in some instances at the ED 16 level. However, the logprobit and one-hit models predict effects of low dosages quite different from those predicted by the logprobit model. The tails of the distributions differ widely even though the central portions are similar (Mantel,

43

1963). Thus, the degree of reduction of the dosage that produces tumors in 1% of subjects necessary to achieve virtual safety (defined as the production of only one tumor in 100 million subjects) differs widely for the three mod­ els, namely, 1/100 for the logprobit model, 1/100,000 for the logistic model, and 1/1,000,000, for the one-hit model (Mantel and Bryan, 1961). In other words, the curves for these other models deviate upward and to the left com­ pared with the linear logprobit relationship. The objective studies of the effects of small dosages discussed in the foregoing section would test all three models simultaneously. It is not necessary to test extremely low dosages but only to learn whether the observed points deviate from the logprobit relationship, and if so, in what direction. For example, in Fig. 1.13 for groups of 1000 the estimated number of mice developing tumors following a dosage of 0.0039 mg per mouse would be 50 for a nothreshold model, 35 for a lognormal distribution, and 18 if a threshold is involved. After a dosage of 0.00195 mg per mouse, the estimates would be 20, 5, and 0.001 (i.e., zero) mice, respectively. Certainly it ought to be possible to dis­ tinguish values of these magnitudes with only a relatively few repetitions of an experiment. Chemical Basis of Thresholds in Dosage-Response Relationships  A great deal is known about the biotrans­ formation of foreign compounds and also about the effects of toxic substances on the otherwise normal metabolism of the body. This knowledge, explains many toxic actions and dosage-response relationships. However, the biochemical basis of the effects of small dosages is poorly explored, just as their dosage-response relationships are poorly studied. It is clear in a general way that thresholds involve dosage but are not necessarily directly proportional to it, and that they are conditioned by the ability of the body to repair some injuries. Sufficiently small doses are without detectable effect. The effects of somewhat larger doses may be harmless in themselves and completely repaired before the next dose is received. This relationship is well illustrated by the action of inhibitors on enzymes when administered at rates that do not cause illness. What is not clear is the identity and relative impor­ tance of mechanisms that do not correspond directly with differences in dosage and, in this sense, may be regarded as qualitative differences. It is often speculated that small doses are biotransformed by normal pathways without tax­ ing them, but that larger doses saturate these pathways, flood others, and thus interfere with endogenous metabo­ lism. Unfortunately, details frequently are lacking, but there are notable exceptions, some of which are discussed in the following paragraphs. Capacity for biotransformation may explain the pres­ ence of a threshold. Furthermore, biotransformation may be one mechanism of repair as is illustrated by the classical

44

example of cyanide poisoning. Prompt metabolism of the cyanide ion to the much less toxic thiocyanate ion serves to prevent a combination of cyanide with cytochrome oxi­ dase (the biochemical lesion in this instance). However, if the lesion already has formed, metabolism of cyanide to thiocyanate helps to establish a gradient that favors release of cyanide from combination with cytochrome oxidase and in this way promotes repair of the biochemical lesion. The tremendous efficiency of the metabolism of small doses of cyanide to thiocyanate explains why we are able to with­ stand the small amounts of cyanide we receive daily from food and other sources. However, above a certain thresh­ old, cyanide is dangerous. In this instance, the limiting factor is not the capacity of the enzyme thiosulfate sulfo­ transferase but the immediate availability of sulfur to form thiocyanate. Moderate doses of cyanide cannot be metabo­ lized efficiently because the sulfur compounds ordinarily used for this purpose are limited in availability. That the limitation of sulfur is, in fact, the reason for the threshold above which cyanide becomes dangerous is demonstrated by the fact that the threshold is moved upward if a suit­ able source of sulfur is furnished. The difference can be measured best not in terms of the threshold itself, but in terms of the LD 50. It was shown by Way et al. (1966) that the LD 50 of potassium cyanide can be shifted from 9 to 33 mg/kg merely by supplying sodium thiosulfate (see Way et al., 1966). Glyoxylate is significantly more toxic than ethylene gly­ col, of which it is a metabolite, and it probably is largely responsible for the toxicity of the parent compound. When the dosage of glyoxylate to monkeys was reduced from 500 to 60 mg/kg, the proportion excreted unchanged was reduced from a maximum of 59% to a maximum of 1.5% and the proportion metabolized to carbon dioxide increased. Thus the kidney, which is specifically susceptible to injury by ethylene glycol and glyoxylate, is protected to a dis­ proportionate degree by the metabolism of low dosages as ­compared with the metabolism of high dosages (McChesney et al., 1972). Liver glutathione (GSH) has a relation to the toxicity of bromobenzene somewhat analogous to that of available sul­ fur to the toxicity of cyanide. There is a close relationship between the covalent binding of halogenated benzenes and their ability to cause necrosis of the liver. However, covalent binding of bromobenzene metabolites to mouse liver protein remains low until a critical dosage of 1.20–2.15 mmol/kg is reached. At dosages of 2.15 and 4.06 mmol/kg (which produces minimal and extensive toxicity, respectively), the rate of covalent binding is not only high, but is over twice what would be predicted by extrapolation of the rates for lower, nontoxic dosages (Reid and Krishna, 1973). Bromobenzene, or especially its epoxide, depletes liver glu­ tathione in the process of forming a mercapturic acid. Little covalent binding of bromobenzene metabolites occurs while the supply of GSH is adequate and mercapturic acid

Hayes’ Handbook of Pesticide Toxicology

is being formed, but considerable covalent binding occurs when 90% of the liver GSH is lost and little mercapturic acid can be formed (Jollow et al., 1974). Anything that reduces liver GSH (even though harmless in itself) makes the liver more susceptible to injury by bromobenzene. Finally, as outlined in Section 1.1.4, bromobenzene is capa­ ble of metabolism by different pathways and the protection is at least partially independent of GSH availability. Thus the metabolism of bromobenzene is complex, but the avail­ able facts all help to explain its disproportionate increase in toxicity above a threshold. As reviewed by Gillette (1973), the toxicity of acet­ aminophen shows a disproportionate increase above a threshold and this relationship depends at least in part on the availability of liver GSH. Another example involves the onset of toxicity at dos­ age levels that exceed the metabolic capacity of the liver. Golberg et al. (1967) found that 2,4,6-tri-tertbutylphenol at a dosage of 0.5 mg/kg/day for 6 days induced hepatic pro­ cessing enzymes, and the efficiency of the enzymes was such that the concentration of the compound in the liver was distinguishable after dosages of 10, 25, and 50 mg/kg/day. However, a dosage of 75 mg/kg/day produced an approxi­ mately eightfold increase in average liver storage, and it was only at this threshold dosage range of 50–75 mg/kg/day that toxicity as indicated by histopathological changes in the liver first appeared. In a study of a series of substituted phenols, it was found that liver processing enzymes invari­ ably were induced by dosages lower than those required to alter the activity of liver microsomal phosphatases or to produce histopathological change (Golberg et al., 1967). Stokinger (1953) reviewed evidence that the tissue distributions of beryllium, silver, iron, and iodine dif­ fer, sometimes greatly, according to dosage. His interest was focused on the serious errors that may be introduced by extrapolating the results of the storage of small tracer doses to the storage of a therapeutic or even toxic dose. Although his interest was in tissue distribution rather than toxicity per se, he noted that one might expect to find a dif­ ferent pattern of toxic manifestations solely because of the different amounts of the toxic agents in various organ sites. Stokinger suggested several mechanisms governing the distribution of the elements listed: (a) dosage-dependent formation of colloidal hydroxides such as those of beryl­ lium, which are then phagocytized by cells of the reticulo­ endothelial system, (b) formation of complexes with serum proteins or other colloids (e.g., complexes of silver), which are then phagocytized, and (c) complex physiological regu­ lation such as that of iodine or of iron. As discussed earlier, nonlinear dosage-dependent dif­ ferences in the toxicity of foreign organic compounds, unlike those of the elements, are likely to depend on other mechanisms, namely, (a) biotransformation and (b) biore­ pair. There may be other mechanisms also, for the subject has been studied inadequately.

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

It has been pointed out that the only beneficial effects of small dosages of toxic substances that now are accepted generally are those that are understood. This uncompro­ mising demand for intellectual justification is admirable. It is clear that the concept of threshold and the observed ben­ eficial effects of at least some small dosages will be more readily accepted if their biochemical basis is elucidated more completely. Studies of dosage-related biochemistry ought to go hand in hand with statistical studies of the clin­ ical effects of small dosages. Discussion  In the exposition of statistical studies involving large groups of animals in Section 1.3.1.3, the ED 01 study of the carcinogen 2-AAF was considered. No biochemical study has explored whether there is any difference between the way the compound is metabolized at dietary levels below 60 ppm and the way it is metabolized at dietary lev­ els of 60 ppm and above that would explain the form of the curves for bladder cancer that we have discussed, Similarly, there are no biochemical studies to explore the possibility of a threshold in the dosage range we have mentioned for liver tumors in mice. One variable that certainly ought to be explored in connection with the liver tumors is the induc­ tion of microsomal enzymes, because it has been suggested that this induction is intimately tied to the tumorigenicity of a number of chlorinated hydrocarbons as well as to that of phenobarbital [World Health Organization (WHO), 1979]. In fact, the same World Health Organization expert com­ mittee recommended specifically that pyrethrins be tested to see whether they increase the incidence of liver tumors in animals, inasmuch as they are known to induce micro­ somal enzymes of the liver. The results of the bioassays were mixed, although largely negative (ATSDR, 2003). WHO (2001) concluded that there is little indication that pyrethroids should be considered carcinogenic, proving the point of Dr. Hayes. In the United States, there is extensive concern among the general population about the safety of chemicals. How much concern would remain if the matter were not continuously inflamed by the media is an open question. Regardless of the source and degree of concern, toxicologists would be in a vastly better position to advise if there were more information on the effects of small dosages. The small number of studies that have been carried out simply is not enough. Only a poor experiment does not raise more questions than it answers. Even those who disagree with the interpretation of the ED 01 study in relation to hepatic tumors could hardly argue that something might not be learned by repeating the study and including dosages at and below the predicted intercepts of the logprobit curves with the control levels. It also would be diffi­ cult to argue that something might not be learned by compar­ ing the dosages of 2-AAF that induce microsomal enzymes with those that increase the incidence of liver tumors in mice. The basic reason for wanting to learn more about the effects of small dosages is scientific and, therefore, intellectual.

45

To be sure, any progress that might be made could have important implications for the regulation of chemicals and for our confidence in that regulation. However, no scientific or practical progress can be made by ignoring the statistical results for control animals or by imagining that studies of chemical or physical carcinogens provide any information about the cause of neoplasia among the controls. In short, no progress can be made by those who attempt to extrapo­ late from animal experiments to humans without obtaining more information than can be measured in the animals. There is an analogy-and a contrast-between the con­ cept of the cumulative lognormal curve in toxicology and Einstein’s proposal for the equivalence of energy and matter (E  mc2). Both concepts were based on theoretical consider­ ations with little or no support from common sense and both were unproved when first proposed. That is where the anal­ ogy ends. Physicists have sought every possible way of test­ ing the validity of the equation E  mc2. They have carried out meticulous experiments, and they have gradually accu­ mulated evidence that has made the equation a cornerstone of modern physics. Toxicologists, on the other hand, have done so little to prove or disprove the theory of the lognor­ mal response of organisms to toxicants that most of the work and the theory have been reviewed in two brief sections. If we toxicologists had been as thorough and energetic as the physicists, we should not be in the strange position of having a theory with no known exception but with so little critical evidence supporting it that few dare to accept it as true. Hayes’ truly visionary words about the need for a the­ ory of toxicology and the dire consequences of claiming exceptions to it are finally coming to fruition in this revised chapter. As postulated in Section 1.1.3 there are no excep­ tions to the c  t concept; there are only incompletely con­ trolled experiments or experiments conducted under less than ideal conditions. It must be emphasized that this can be seen only under worst case exposure conditions (contin­ uous exposure) or when an effect is essentially irreversible during the observation period. Somehow we need to rec­ ognize the futility of continuing to conduct toxicological experiments in the traditional way. Otherwise, we are just producing some more of the data that raise more questions than provide answers. It is entirely meaningless to continue examining the myriad of microscopic variables without the guiding constraint of laws of the macroscopic variables, which are the dose (concentration at the site of action), the various timescales and the effect. One more issue needs clarification for the sake of mak­ ing this section complete. The principles underlying bene­ ficial and detrimental effects are the same and as such this distinction is highly anthropocentric and not scientific. For example, low doses of many bacteriostatic agents promote bacterial growth (which is undesirable from an anthropo­ centric view, but desirable from a bacterial point of view), whereas high doses kill the bacteria (which is desirable from the anthropocentric viewpoint, but not so from a bacterial

46

point of view). Another example is the issue of contracep­ tives. For a couple not wishing to have children, oral contra­ ception is a beneficial effect (desirable). For another couple, yearning for children without fulfillment, the presence of naturally occurring contraceptives in the diet would be an adverse effect (undesirable). Toxicology would be better served and would remain a more credible science if such value-laden terms were avoided. An effect is a dose- and time-dependent action of a chemical on an organism charac­ terized by one or more dose- and time-dependent responses. Low-dose effects are most often beneficial to an organism whereas high-dose effects are for the most part adverse (toxic) effects. We need to remember though that selective toxicity (e.g., antibiotics, pesticides, cancer therapy) can be highly desirable, because of the perceived benefit.

1.3.7.5  Geometric Mean Francis Galton (1879) pointed out that, in many vital phe­ nomena, equal intervals of effect are produced by logarith­ mic intervals of stimulus. He used as a specific example Fechner’s law, which in its simplest form states that sensa­ tion is proportional to the log of stimulus. Galton empha­ sized that, for such phenomena, the true mean is the geometric one. In the geometric series 1, 2, 4, 8, 16, 32, .., the geometric mean of 4 and 16 is 8 (i.e., 4 : 8  8 : 16) and not 10 (i.e., not (4  l6)/2). Use of the geometric mean where appropriate avoids the consequences of assuming that errors in excess or in deficiency of the truth are equally probable. To show how absurd or misleading this assump­ tion can be, Galton recalled that, because there are giants more than twice as tall as the mean height of their race, the assumption “implies the possibility of the existence of dwarfs whose stature is less than nothing at all.” In his brief rational paper, Galton introduced a more technical mathematical study by Donald McAlister (1879) entitled “The law of the geometric mean.” This law has not received the attention or use it deserves. It is appro­ priate for calculating the average storage of a compound in a population or the average time of death in a series of ­animals all dosed in the same way. On the other hand, few of the published arithmetic means are so much in error that they ought to be discarded. As Galton (1879) noted, the difference between the arithmetic and the geometric mean is small if the range of the values averaged is narrow.

1.3.7.6  Reproducibility of Results Ideally, the results of any particular measurement ought to be reproducible in the same laboratory or from one laboratory to another. This becomes especially important when the numer­ ical results may be used as guides for diagnosis and therapy, or when any results may be used to determine whether a compound does or does not satisfy legal criteria (e.g., criteria

Hayes’ Handbook of Pesticide Toxicology

of registration or residue tolerances). However, results can be meaningful and important even when it is impossible to standardize the conditions to the point that control values are statistically identical from one trial to another. An astonishing proportion of biological and biochemi­ cal studies are recognized as valuable contributions if the results for each experimental group show a clear-cut rela­ tion to the results for the corresponding control in the same experiment. Of course, no study can be considered con­ firmed until the relationships demonstrated in the initial experiment have been redemonstrated in the same labora­ tory or, even better, in different laboratories. Whereas all scientific procedures are examined from the standpoint of reproducibility within an experiment, only a few toxicological methods have been examined thoroughly for reproducibility in a broader sense. In these studies, it sel­ dom has been possible to identify all of the causes of varia­ tion. In animal experiments, some of the variables discussed in Section 1.4 may be detected (Weil and Wright, 1967). Probably the most important single factor in determin­ ing reproducibility is the objectivity of the end point. In a study of the oral LD 50, for which the end point is clearcut, different protocols in use in well-established labora­ tories produced results that differed so little that choice of one or the other would not change the interpretation of the relative hazard of any particular compound. Specifically, the highest and lowest LD 50 values for each of 10 com­ pounds as determined in eight laboratories by various pro­ tocols differed by factors ranging from 1.30 to 5.48. The degree of variation was less, but not statistically less, when each laboratory used a reference protocol and a reference stock of rats as compared with (a) reference protocol and rats commonly used in the laboratory or (b) both proto­ col and rats commonly used in the laboratory (Weil and Wright, 1967). Far greater differences were found in a study of intral­ aboratory and interlaboratory variability in the results of eye and skin irritation tests, for which the end points are subjec­ tive. Although other factors were involved, it was concluded that the main factor contributing to variability was difficulty in reading the reactions. Although numerical factors of dif­ ference (between highest and lowest values) could not be assigned, some of the differences obviously were very great. The majority of laboratories performed the tests competently and reproducibly; however, others were far afield. Some materials were rated the most irritating by some laboratories and rated the least irritating by others. Some of the labora­ tories that were most out of line were industrial and some were governmental. Therefore, restricting testing to any one type of laboratory would not solve the problem. In fact, it was concluded that the tests that had been in general use for 20 years were no longer dependable ways of classifying a material as an irritant or a nonirritant. It was suggested that modification of the tests themselves would not be helpful but

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

that careful reeducation of those who perform the tests would be required if any improvement were to be made (Weil and Scala, 1971). The Council of the Society of Toxicology sup­ ported this emphasis on training and a lack of emphasis on rigid standardization of protocols (Hayes et al., 1971). One factor that may contribute to the failure of a labora­ tory to agree with the majority of others in a particular test is unfamiliarity with the test. In a study of the reproducibil­ ity of measurements of blood lead, it was noted that some of the laboratories ordinarily had occasion to use the test only a few times per year (Keppler et al., 1970). Here is not only one explanation for poor performance but an indication that the study may not have reflected the accuracy of experi­ enced laboratories. Reeducation would be most efficient if it could be provided when needed, that is, just before an infre­ quently used test is required. However, if only a few tests are to be run, it probably would be more efficient to refer them to another laboratory than to arrange training.

1.3.7.7  Abnormal Values in Control Groups It sometimes occurs that a statistically significant dif­ ference between an experimental group and its control depends on an abnormality of the control and not on any deviation in the experimental group. This is an important reason changes, to be indicative of a deleterious effect, must be produced that are dosage-related and illustrate a trend away from the norm for the population under study (Task Force of Past Presidents, 1982; Weil et al., 1969).

1.4  Dosage-response relationships in different kinds of toxicity or change 1.4.1  Toxicity (Sensu Stricto) All people with toxicological or medical training are aware that toxicity in the restricted sense corresponds to dosage for any particular compound. However, use of the various procedures described in Section 1.2 for measuring dosageresponse relationships is restricted all too often to this lim­ ited kind of toxicity. Toxicity in both the strict and broad sense consists of ill­ ness or death. It sometimes is implied, without toxicological or logical basis, that specifying the kind of illness involves some change in underlying principle over and above the restrictions imposed by the specification. Neurotoxicity is restricted to the nervous system and teratology to the embryonic stages, but the broad principles of toxicology remain unchanged. The less common or less familiar a phenomenon is, the more likely that its relationship to an actual or supposed

47

etiology will be viewed qualitatively rather than quan­ tita­tively. Phenomena often viewed in this way include neuro­toxicity, teratogenesis, carcinogenesis, mutagenesis, hypersensitivity, and storage, as well as adaptive response of microsomal enzymes.

1.4.2  Neurotoxicity Neurotoxicity is the delayed but persistent paralysis (poly­ neuropathy) caused by certain organic phosphorus com­ pounds as well as a few other toxicants, deficiency diseases, and infections. The classical example is “jake leg” paraly­ sis caused by triorthocresyl phosphate (Smith et al., 1930). Most studies of neurotoxicity associated with organic phos­ phorus compounds have attempted to learn which molecu­ lar configurations are capable of producing the phenomenon and which are not. When active compounds were investi­ gated qualitatively, it was found that a sufficiently low dos­ age was tolerated, and progressively larger dosages increased the frequency and severity, and often reduced the latency of neurotoxicity (Aldridge and Barnes, 1961; Cavanagh et al., 1961; Davies et al., 1960; Siegel et al., 1965). Hayes (1991) points out the time-dependence of neuro­ toxicity, but also that clear-cut quantitative relationships have not been found, perhaps because they have not been sought. Part of the problem why such studies have not been con­ ducted is a lack of conceptual framework for the experimen­ tal design in combination with formidable methodological difficulties. Neurotoxicity is most often irreversible to some extent, although adaptation to neuronal damage is possible as well as repair of the injury in some instances. Thus, in many instances when the underlying dynamic processes are slower than the kinetics (elimination) of the causative agent there will be two or more rate-limiting steps in the recovery pro­ cess from an insult. Whichever will be the rate-­determining step could be elucidated by a careful time course study of the recovery and curve stripping to find the respective halflives of adaptation, repair, and reversibility. Such experi­ ments have not been conducted to our knowledge. Other types of neurotoxicity, such as loss of the righting reflex or unintentional anesthesia (which are highly reversible), have (a) kinetic process(es) as the rate-determining (-limiting) step(s) in their action. Under conditions of continuous expo­ sure (inhalation) such effects have been often shown to obey Haber’s rule of c  t  k (Flury and Wirth, 1934).

1.4.3  Teratogenesis Although most research in teratogenesis has been cen­ tered on the nature of the phenomenon itself as well as the biological factors, which influence it, a few quantita­ tive studies have been made. These studies illustrate that for any given compound and experimental situation there is a dependable relationship between dosage and effect

Hayes’ Handbook of Pesticide Toxicology

48

(Murphy, 1965; Wilson, 1964). It has been pointed out that there is no way to exclude that any given compound may be teratogenic to some species under certain conditions (Bough et al., 1963; Karnofsky, 1965). Steep dosage-response curves for teratogenic action such as those shown in Fig. 1.14 are not uncommon and, in fact, appear to be the rule. Agents can be tolerated in low dos­ age without any recognizable effect on development or via­ bility, but most of them that have detectable teratogenicity rather quickly become lethal to all embryos at higher dos­ ages. Between these ranges of normality and lethality, there exists a narrow zone of dosages in which variable numbers of embryos survive with varying degrees of teratogenic involvement. A sharp rise of the dosage-response curve is also characteristic of the teratogenic action of X-radiation (Wilson, 1964). From the theoretical point of view, the critical tim­ escale in teratogenesis is that of organogenesis, which is such a narrow window in time that it will be difficult if not impossible to determine time-dependence of a teratogenic effect (within that window) experimentally. However, teratogenicity occurs at dynamic steady state because no recovery from it is possible after the narrow window in time passes. In agreement with previous considerations this should yield a very steep dose response because once again nature provides nearly “ideal conditions” for teratogenic experiments. Indeed, Fig. 1.14 demonstrates that the dose-

90 80

responses for diphenylhydantoininduced orofacial and skeletal anomalies occur with a slope of less than 2. The importance of both kinetic and dynamic considerations in the manifestation of teratogenicity of retinoic acid agonists has been highlighted in Arafa et al. (2000).

1.4.4  Carcinogenesis Strong carcinogens demonstrate a striking dosage-response relationship whether expressed on the usual basis of incidence versus dosage (Figs. 1.13 and 1.15) or on the basis of logtime versus logdosage (Fig. 1.11). Thorough reviews of the matter have been written by Druckrey (1967) and by Shabad (1971). In many instances the maximal tolerated dosage of weaker carcinogens is required to reveal carcinogenesis with statisti­ cal dependability. From a data base of 52 studies, Haseman (1985) tabulated examples from which he concluded that over two-thirds of the positive results would have been missed if only half the maximal tolerated dosages of the different com­ pounds had been studied. However, so much research on carcinogenesis has been centered on the phenomenon itself and so little attention has been given to quantitation of the actions of chemicals that even some experts in the field seem unfamiliar with the fact that chemical carcinogenesis follows clear-cut dosage-response relationships. There is no doubt that only a small dosage of certain compounds is necessary to increase the incidence of cancer in susceptible animals. For example, as shown in Table 1.5, the ED 50 for carcinogenesis following a single ­subcutaneous injection of three of the classical laboratory carcinogens in mice ranges from 0.76 to 4.6 mg/kg. The naturally occurring carcinogen aflatoxin also is effective when administered sub­ cutaneously at a total dosage of about 1.8 mg/kg (Dickens

60 50

100

40 30

80 Incidence (%)

Anomalies (%)

70

20 10

60 40

5

20

2

0 0.0001

1 20

30

50 70 100 Dosage (mg/kg)

200

FIGURE 1.14 Incidence of orofacial anomalies (O) and skeletal anomalies (•) in fetal mice whose mothers received different dosages of diphenylhydantoin by intraperitoneal injection on gestation days 11, 12, and 13. Data from Harbison and Becker (1970).

0.001

0.01

0.1

1.0

10

Dose (mg) FIGURE 1.15 Responses of mice to graded single doses of each of three polycyclic aromatic hydrocarbos carcinogens dissolved in tricaprylin and injected subcutaneously: methylcholanthrene (O); dibenzanthracene (Δ); 3,4-benzo[a]pyrene (䊐). Approximately 20 animals per dose of each compound. From National Academy of Sciences (1960) by permission of the Academy.

Chapter | 1

Dose and Time Determining, and Other Factors Influencing, Toxicity

49

TABLE 1.5 Toxicity of One Dose of Selected Materials Material (route)a

Species

Sex

Dosage (mg/kg)

Effect

Reference

α-fraction (iv)

Mouse

—

0.000,000,27

LD 50

Dasgupta et al. (1966)

Unfractionated (ip)

Mouse

—

0.000,001,4

LD 50

Lamanna and Carr (1967)

Same (po)

Mouse

—

0.001,4

LD 50

Lamanna and Carr (1967)

Same (po)

Human

—

0.000,014

LD 50

Schantz and Sugiyama (1974)

O-ethylmethyl-S-phosphorylthiocholine iodide (ip)

Human

—

0.03

LD 50

Holmstedt (1959)

N, N’-Di-n-butylphosphorodismine fluoride (im)

Chicken

F

2.0

LD 50

Davies et al. (1966)

Same (im)

Chicken

F

0.05

parab

Davies et al. (1966)

Botulinal toxin A

Dibenzanthracene (sc) Methylcholanthrene (sc)

Mouse Mouse

M M

0.76 0.96

c

Bryan and Shimkin (1943)

c

Bryan and Shimkin (1943)

c

ED 50 ED 50

Benzo[ a]pyrene (sc)

Mouse

M

4.6

ED 50

Bryan and Shimkin (1943)

Aldicarb

Rat

M

0.8

LD 50

Gaines (1969)

Tetraethylpyrophosphate

Rat

M

0.8

LD 50

Gaines (1969)

Parathion

Rat

M

13

LD 50

Gaines (1960)

Endrin

Rat

M

17.8

LD 50

Gaines (1960)

Arsenic trioxide

Rat

M

72

LD 50

Gaines (1968)d

Nicotine sulfate

Rat

F

83

LD 50

Gaines (1960)

DDT

Rat

M

113

LD 50

Gaines (1960)

Pyrethrum

Rat

M

470

LD 50

Gaines (1968)d

Acetylsalicylic acid

Rat

—

1360

LD 50

Eagle and Carlson (1950)

Malathion

Rat

M

1375

LD 50

Gaines (1960)

Sodium chloride

Rat

M

3550

LD 50

Gaines (1968)d

Difenphos

Rat

M

8600

LD 50

Gaines et al. (1967)

a

Doses are oral unless otherwise indicated; iv, intravenous; ip, intraperitoneal; po, per os; im, intramuscular; sc, subcutaneous. Para, paralysis. c Carcinogenesis. d T: B. Gaines; personal communication to W. J. Hayes, Jr. (1968). b

et al., 1966), but its danger by the oral route is more impor­ tant. Dietary intake of aflatoxin B1 for only 2 weeks at a total dosage of about 2.6 mg/kg produces carcinoma in male rats and a lower daily intake for a longer period also pro­ duces cancer when the total dosage is less than 0.5 mg/kg (Wogan and Newberne, 1967). Unfortunately, little or no attention has been given to the dosage-response relationships of weak carcinogens. Problems of the quantitative study of the effects of small dosages discussed in Section 1.3.7.4 are relevant to carcinogens as well as other toxicants. In fact, Fig. 1.13 is based on a study of 20-methylcholanthrene, and Fig. 1.15 is based on a study of three carcinogens.

There is evidence, sometimes of a very tenuous nature, that some pesticides are tumorigens if not carcinogens. Hayes (1991) recognized that carcinogenesis is one of the most distorted issues of science mainly because of the enormous societal concern about cancer, which has become one of the major causes of old-age-related death. Physicians have been cognizant of the capability of chemi­ cal and physical agents to induce cancer at least since Sir Parcifal Pott’s observation of scrotal cancer in chimney sweeps. This observation was confirmed experimentally in controlled studies in the early 20th century (Yamagawa and Ichikawa, 1915). Therefore, chemical-induced cancer was considered early on as one end point of toxicity. Initially,

50

potent carcinogens were studied for which the ­carcinogenic dose-responses were to the left of nonspecific toxicity and/or old-age-related death. Therefore, complete and steep dose-responses were obtained as shown by Fig. 1.15. The slopes of these dose-responses are somewhat distorted, because of plotting dose rate (daily dose) instead of dose (­cumulative dose) vs. effect. When plotted on the appropri­ ate dose scale the carcinogenic dose responses are as steep as any other dose-responses discussed thus far (Rozman et al., 1996). When toxicologists started studying less potent carcinogens the dose-responses became truncated by the 2-year terminal sacrifice or the natural life span of the experimental animals. Introduction of the MTD reduced carcinogenesis to a largely qualitative yes-or-no-type phe­ nomenon determined by a statistical comparison of treated animals to controls. The fact that at MTD/2 about 50% of the chemicals have shown no statistical difference to controls indicates, in agreement with theory, that most of the recently conducted bioassays targeted weak to very weak carcino­ gens for which the carcinogenic dose response coincides with the dose-response of nonspecific toxicity. As is the case with other end points of toxicity, there are no shallow cancer dose-responses, there are truncated cancer dose-responses, incorrectly plotted cancer dose-responses, and there are incompletely controlled cancer studies measuring at the same time toxicity and recovery when the kinetic or dynamic half-life of a carcinogen is very short and dosing occurs once or twice (feeding) a day. It is unfortunate that toxicologists surrendered their knowledge and understanding of toxic phe­ nomena to the modelers and to statistics, which describe that which is unknown or unknowable. The price paid has been a stagnation of the development of the theory of toxicology.

1.4.5  Mutagenesis Nearly all studies of chemical mutagenesis have been concerned with identifying mechanisms of action or with learning whether selected chemicals can or cannot cause some mutagenic effect in the system under study. Little attention has been paid to dosage-response relationships. However, when such a relationship is sought it has always been found. Examples from his own work and that of oth­ ers on the induction of phage were given by Heinemann (1971). Dominant lethal mutations produced in insects and mammals by many alkylating agents and some other com­ pounds regularly show dosage-response relationships. The toxicity of some of these compounds considered for use as insect chemosterilants has been reviewed (Hayes, 1964, 1968). Mutagenicity is one of the key lessons of modern toxicology. Early on, it was hailed as an inexpensive way to predict carcinogenicity. Later, the claim was reduced to at least predicting DNA-reactivity. Neither hope was ful­ filled and this widely used assay is on its way to the role of an ignominious prediction never materialized.

Hayes’ Handbook of Pesticide Toxicology

1.4.6  Hypersensitivity and Allergy The extreme sensitivity of some people to certain sub­ stances is illustrated by the fact that anaphylactic reactions to penicillin have been produced during skin testing with as little as 10 units of the drug (Mayer et al., 1953). The fact that some people are allergic and others are either not allergic or even highly resistant tends to obscure the fact that the various forms of hypersensitivity are dosagerelated within a homogeneous population. Individuals who suffer from allergy often find that a reduction of dosage will lead to clinical improvement. For example, people who are sensitive to pollen often get some relief by closing most of the air inlets of their homes and placing filters on the remain­ ing ones, even though this procedure does not eliminate their exposure to pollen but merely reduces it. Under experimen­ tal conditions in which the susceptibility of animals was made uniform by passive transfer, the onset of anaphylaxis was directly related to the doses of antigen (Pruzansky et al., 1959). Human leukocytes isolated from ragweed­sensitive donors release histamine at rates determined by the concentrations of purified antigen derived from the pollen (Lichtenstein and Osler, 1964). Other forms of hypersensi­ tivity may be dosage-related also. For example, blood dys­ crasias, especially aplastic anemia, are a recognized hazard of the otherwise valuable drug chloramphenicol. Hodgkinson (1954) showed that these dangerous side effects occurred predominately in cases in which the drug had been adminis­ tered at a rate significantly higher than usual. The fact that even hypersensitivity is often dosage-related emphasizes the importance of searching for suspected but undemonstrated dangers of a particular compound among people whose exposure is most intensive and prolonged. Hypersensitivity and allergy are special cases of toxic responses, which occur in subjects that are not part of the normal distribution in terms of this particular and possi­ bly of some other responses. Expressed differently, once an individual has been sensitized, it is no longer the same subject as before. Rather, the sensitized individual belongs to the normal distribution of a sensitized population with its own dose- and time-responses. It will be very difficult to sort out dose- and time-responses in such individuals, because the sensitization occurs in an individual who is still part of the normal distribution of the general popula­ tion while being sensitized but becomes part of another normal distribution thereafter.

1.4.7  Induction of Enzymes Microsomal enzymes, offer some explanation for a num­ ber of otherwise obscure facts in toxicology. It is generally admitted that the net effect of these enzymes is adaptive for the organisms, but it has been suggested that stimulation of enzymes by one chemical will lead to greater injury to the organism when faced with some other challenge. Regardless

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

of the final toxicological evaluation of the process of induc­ tion, it is clear from work already completed that there are orderly relationships between dosage and response for com­ pounds that stimulate microsomal enzymes. In fact, one of the very early papers by Conney et al. (1956) demonstrated very clear dosage-response relationships for demethylase and DAB-reductase following injection of 3-methylcholan­ threne. The same paper demonstrated a similarly clear rela­ tionship for inhibition by ethionine. Apparently the first such studies of pesticides as inducers of microsomal enzymes were those of Hart and his colleagues regarding chlordane and DDT. No obvious ­ dosage-response relationships were found with either single or multiple doses of chlordane (Hart and Fouts, 1963). Some indication of a dosage- response relationship was evidence from tabu­ lar values for DDT, but the relationship apparently was not discussed (Hart and Fouts, 1963) or discussed only briefly (Hart and Fouts, 1965) in connection with these early studies. Kinoshita et al. (1966) first demonstrated clearly a dosagerelated effect of DDT and toxaphene on enzyme induction. The result has been confirmed in connection with DDT (Gillett, 1968; Hoffman et al., 1970) and other compounds (Gielen and Nebert, 1971; Hoffman et al., 1968). Sufficiently small dosages produce no detectable effect on enzyme activity. There is considerable evidence that the threshold dosage for enzyme induction corresponds to the upper limit of intake that can be metabolized by the unstimulated liver (Hoffman et al., 1970). The threshold dosage of DDT for induction of various microsomal enzymes in the rat has been estimated at about 0.05 mg/kg/day (i.e., a dietary level of 1 ppm) (Kinoshita et al., 1966) or 0.5 mg/kg/day (Schwabe and Wendling, 1967). Datta and Nelson (1968) found that a dietary level of 4 ppm (about 0.2 mg/kg/day) induced enzymes. Gillett (1968) found the threshold to be 0.125 mg/kg/day. Street et al. (1969) estimated the threshold at 0.05 mg/kg/day. The different estimates are not necessarily inconsistent, because they depend on different test systems. In any event, the lowest estimate (0.05 mg/kg/day) is only 0.2 times that known to be effective in humans (Laws et al., 1967; Poland et al., 1970) whereas it is 50 times greater than the average dietary intake of all DDT-related materi­ als by a 16- to 19-year-old man during the mid l960s, that is, 0.0009 mg/kg/day (Duggan, 1968). The enzyme-inducing dosage of DDT (0.5 mg/kg/day) used by Schwabe and Wendling (1967) led in 14 days to a storage level of 10 ppm in the adipose tissue of rats. The dosage of 0.2 mg/kg/day used by Datta and Nelson pro­ duced in 20 weeks a storage of 39 and 76 ppm in the adi­ pose tissue of male and female rats, respectively. Twelve weeks after dietary feeding of DDT was stopped, the stor­ age levels of DDT-related materials had fallen to 11 and 21 ppm in males and females, respectively, compared with 6 and 9 ppm, respectively, in the controls. The rats previously fed DDT still showed some induction of liver enzymes 12 weeks after dosing was stopped. Neither the

51

rats described by Schwabe and Wendling (1967) nor those described by Datta and Nelson showed a steady state of DDT storage when their values were between 10 and 21 ppm. It is, therefore, open to serious question whether these storage values are at all comparable with those found in people in the general population. The order of Sections 1.4.8 and 1.4.7 has been reversed in this edition also, because enzyme induction is part of the dynamics of a chemical (what does the chemical do to the organism?) and as such belongs to the same category as neurotoxicity, cancer, etc. Metabolism and storage, on the other hand, are part of the kinetics of a chemical (What does the organism do to the chemical?). Enzyme induction is for the most part a transient adap­ tive response of an organism which promotes biotrans­ formation of the causative agent and thereby contributes together with excretion to detoxification (recovery by kinetics). In rare instances, enzyme induction will lead to metabolic activation, namely, making the chemical itself or other compounds more toxic to the host (Parkinson, 1996). Enzyme induction is a highly reversible phenomenon and if the half-life of the causative agent is short, there will be few if any adverse consequences to the host unless exposure to the chemical is continuous. The other limit­ ing condition is represented by chemicals having very long kinetic half-lives, in which case even after a 90-day off-dose period there is virtually no reversibility of induction observ­ able (Viluksela et al., 1997). This is the kinetic equivalent of an essentially permanently altered organism as discussed for hypersensitivity, which in contrast to this has dynamic causes. As long as the induction persists, the individual belongs to a different population in terms of normal dis­ tribution. Thus, he or she may be more or less sensitive to other toxic agents depending on whether metabolic activa­ tion or deactivation (detoxification) is the rate-­determining step. For example, induction of enzymes metabolizing acet­ aminophen to its toxic metabolite aggravates its hepatotoxic­ ity, but enzyme-inducing doses of TCDD reduce mammary tumors highly significantly below controls (Rozman et al., 1993, 1996; Rozman et al., 2005). Enzyme induction per se is not a toxic effect; it is just an effect. It depends on the consequences whether or not enzyme induction will lead to decreased or increased toxicity or remains inconsequential for the host. A lack of conceptualization has resulted in the loss of a great many potentially superior drugs.

1.4.8  Metabolism and Storage The relationship between equilibrium storage and daily dos­ age for DDT in the human, rat, rhesus monkey, dog, and tur­ key are shown in Fig 1.16. It is clear that equilibrium storage corresponds to daily dosage in all species studied. However, the details of this relationship differ according to species and, at least in the rat, according to sex and dosage level. Specifically, storage is the same in male and female rats up

Hayes’ Handbook of Pesticide Toxicology

52

to a dosage of about 0.02 mg/kg/day, but above this level storage is greater in females. This either represents an exper­ imental artifact or some other error, because male rats must have lower storage levels than females, because of greater growth delution (faster growth). Although species differ­ ences are to be expected, the pattern reported for the dog is remarkably different from the patterns for other species and ought to be reexplored, especially at lower dosage levels. In humans, some elements and compounds are stored in progressively greater concentration with increasing age. This requires further study to determine the cause in each instance; it is not a reason to doubt the general prin­ ciple of equilibrium. Possible causes include (a) excretion so slow that equilibrium is not achieved during the inter­ val involved, (b) a combination of very slow excretion and decreasing dosage so that storage in older people still reflects their higher dosage before the younger people were born, and (c) progressive decline in the ability to metabo­ lize or excrete the material based either on a specific injury by the toxicant (as in the case of radium) or on age per se. Storage is part of the kinetic phenomenon of distribu­ tion, and metabolism is part of the kinetic process of elimi­ nation; and as such both are distinctly different from the dynamic part of the decision tree (Section 1.1.1) repre­ sented by all other subsections of Section 1.4. Storage is due to the fact that one or several body com­ partments with slow blood perfusion have high affinity for a given chemical, which leads to its redistribution pro­ vided that its distribution into these compartments is faster

than its elimination. Thus, the timescale of redistribution depends on the ratio of the distribution and elimination half-lives of a chemical. For example, the initial ­distribution of lead results in very high liver concentrations. The very slow excretion of lead allows for virtually complete redis­ tribution into bone, which is a compartment extremely slow to equilibrate with the central compartment. If the half-life of lead could be decreased to a few minutes, there would be little storage in or redistribution into the bone matrix at all. Data of Fig. 1.16 indicate a linear relationship between dose rate and storage in adipose tissue of DDT in differ­ ent species, whereby the slopes are apparently species­dependent. There is no doubt that during the early phase of a subchronic-chronic study there will be linear accumula­ tion of a compound that has high affinity for a tissue com­ bined with slow excretion. DDT preferentially accumulates in fat (Fig. 1.16), whereas TCDD accumulates to about the same extent in adipose tissue as in liver (Weber et al., 1993). However, after 3.32 distribution half-lives 90% and after 6.64 half-lives 99% of maximum storage will be reached for either compound, implying lack of linearity in the storage of chemicals beyond the initial stage. Different slopes in differ­ ent species and genders are related to different body compo­ sition, differential metabolism, and vastly different growth rates of the species or gender discussed, which – if studied carefully – would account for all the differences. Metabolism or biotransformation of chemicals is clearly an important aspect of their kinetics, but the role of bio­ transformation in toxicology is overrated. Biotransformation

10,000

DDT in fat (ppm)

1000

100

10

1 0.001

0.01

0.1

1

10

100

DDT dosage (mg/kg/day)

Figure 1.16  Storage of DDT in the adipose tissue of human (…), rat (O — O, female; •—•, male), rhesus monkey (—), dog (– · –), and turkey (– – –). The curves have not been extrapolated beyond the dosage levels studied.

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

amounts to altering a chemical into a different chemical moiety, which is one form of elimination. It is a removal of the metabolite from the shifting equilibrium of the par­ ent compound, resulting in the metabolite taking on its own kinetic/dynamic life. Most metabolites are more water soluble (particularly phase II products except a few acety­ lated and methylated compounds) than the parent chemical and hence are excreted more rapidly. After chronic admin­ istration, their steady state concentration is often negli­ gible compared to the parent compound. This amounts to kinetic recovery from intoxication. If a metabolite formed is of comparable or higher intrinsic toxicity (agonists) than the parent compound then the question of additivity must be entertained, provided that the parent compound and metabolite(s) act by the same mechanism. If they act inde­ pendently (different mechanism of action) the more toxic one(s) will dominate the manifestation of toxicity. It is only in very rare instances that in such a constellation we have the appearance of a synergistic interaction, which has either kinetic or dynamic reasons or both in combination.

1.5  Factors influencing toxicity of any kind Although dosage and time are the main factors determin­ ing whether or not a particular chemical will produce a given effect, there are other factors that influence response. Factors of a biological nature include route of exposure, species and individual differences, sex, age, nutrition, and disease. Physiochemical factors include temperature and other environmental variables, and of course, the sched­ ule and duration of dosage and the formulation in which the chemical is administered. There is no theoretical and frequently little practical limit to the range of dosages that may be explored experimentally, and there is frequently little practical limit to the range of human dosages that may be encountered at least occasionally. Thus, at one end of the spectrum it may be possible to find unexposed populations and at the other end of the spectrum to find an occasional person who is killed by accidental inges­ tion of a single large dose. Compared with dosage (and time), the other factors that influence response to a particu­ lar chemical have been subject to less quantitative study. However, some have been studied, and it can be said, for example, that pesticides are, on the average, a little over four times more toxic by the oral route than by the dermal route. Factors other than dosage (and time) are important only in special circumstances. Thus, altitude (atmospheric pressure) is unimportant for toxicology in the parts of the world where most people live, but becomes progressively more important in connection with some compounds as altitude increases from 5000 feet and progressively greater strain is placed on cardiorespiratory function. Certain fac­ tors are almost universally relevant in toxicology and may

53

affect humans, whereas other factors (such as the details of caging) apply directly to animal experiments only. The response of liver microsomal enzymes to dos­ age was mentioned in Section 1.4.7. Differences in these enzymes or in their ability to be induced explain many differences in susceptibility to poisoning associated with interaction of compounds, differences in species, sex, age, and perhaps other factors. In addition to differences in metabolizing enzymes, changes in membranes, ionization, protein binding, and bile flow in some instances may explain observed interac­ tions of compounds or differences of susceptibility to poi­ sons associated with age, diet, and the like. Reviews of the effects of one or more factors provide details and references beyond the scope of this chapter; they include articles by Clough (1982), Fortmeyer (1982), Everett (1984), and Rao (1986). Hayes (1991) identified the four independent variables (compound, subject, dose, and time) of toxicity (effect) and a number of circumstance- or experiment-dependent variables (sex, age, nutrition, disease, temperature, etc.). Of the four fundamental variables of toxicity, compound and ­ subject are implicit variables, because in the absence of one, the other, or both, there is no toxicity. Didactically it would be more advantageous to keep compound and subject (Sections 1.5.2, 1.5.3, 1.5.7, and 1.5.8, 1.5.9, 1.5.10, etc.) together because they determine the qualitative aspects of toxicity (potential spectrum of effects), whereas dose and time pro­ vide the quantitative framework for toxicity (Sections 1.5.1, 1.5.4, 1.5.5, and 1.5.6). It was decided, however, to leave the sequence of sections and subsections intact until perhaps a later edition, to retain the historical perspective, because rearranging the sequence would not alter the issues involved.

1.5.1  Dosage Control of dosage is the basis of almost all safety assess­ ment in the use of chemicals. This rule applies not only to compounds of relatively high toxicity, but also to com­ pounds of low toxicity, including those necessary to life. Babies have been killed by putting salt in their formula in place of sugar (Finberg et al., 1963), and it is said that the ancient Chinese carried out executions using water as a toxicant. On the other hand, all of us tolerate traces of arse­ nic, lead, and mercury (Monier-Williams, 1949), which are naturally occurring elements widely distributed in food and water. They are found in marine fish and in undeveloped areas where they have no use in industry or as pesticides. A sufficiently large dosage of an ordinarily harmless material is fatal. On the other hand, a sufficiently small dosage of the most virulent poison is without effect. For every compound, dosage can make the difference between health and death; in this sense the importance of this factor is infinite (see clarification by Rozman and Doull below).

54

Although age, nutrition, and perhaps other factors may be independent determinants of toxicity in animals of the same strain and sex (Sections 1.5.10 and 1.5.11), it is aston­ ishing how infrequently the effective dosage for small and large (mainly juvenile and mature) animals of the same strain can be distinguished statistically. This conclusion is consistent with the results of a study of botulinum toxin (Lamanna and Hart, 1968), which was certainly the most thorough investigation of the relationship between body size and effective dosage. Even though the extreme affinity of botulinum toxin for its receptor is unique, it still follows that strain and species differences, which involve many com­ pounds and often are substantial, cannot be explained by differences of size per se. Similarly the striking difference of the susceptibility of young and old rats to 1-naphthalenyl thiourea (ANTU) cannot be explained by their size. Dosageresponse relationships that are truly different are empha­ sized, not hidden, when expressed in terms of body weight. Dose and time are inexorably bound together in the c  t relationship above the c  t lifetime threshold if exposure is continuous to manifestation of effect. If all timescales are kept constant then the effect becomes solely dose dependent (see Section 1.1.4). Such “pure” dose responses can be par­ ticularly frequently observed in in vitro experiments with constant incubation time. When the dose is kept constant and one timescale is allowed to vary, “pure” time responses emerge (see Section 1.1.4). Therefore, “pure” dose-andtime responses are limiting cases of the general constella­ tion when toxicity is both dose-and time-dependent. The experiments of Lamanna and Hart (1968) were conducted under isoeffective conditions, because LD 50s were compared. For most of the substances tested a linear relationship was found between log(LD 50) and log(body weight). Geyer et al. (1990, 1993) also found a linear rela­ tionship between log(LD 50) and log(total body fat content) for TCDD among more than 20 species and strains of mam­ mals. They argued that total body fat content was a surro­ gate measure of time because “fatter” animals intoxicated by TCDD lived longer than their leaner counterparts. It could be argued that, like body fat for lipid-soluble compounds, body weight is a good surrogate of time for water-soluble substances. Therefore, it is likely that the good correlation for all but one of the chemicals investigated by Lamanna and Hart (1968) reflects Haber’s rule in its logarithmic form. Of 16 compounds tested, only one appeared to deviate from lin­ earity, which was most likely due to an unidentified variable related to the fuzziness of the end point of measurement. All 15 compounds obeying linearity had short time to death and appeared to cause death by a similar mechanism (neuro­ toxicity) of action. The one anomaly (ANTU) caused lung edema, the development of which can take longer than one day (limit of the observation period). Therefore, almost cer­ tainly supralethal doses must have been used to cause death within one day, which would lead to departure from linearity if deaths due to other causes started occurring concurrently

Hayes’ Handbook of Pesticide Toxicology

(Section 1.3.3.1). This is a very good example to illustrate what a theory-and the lack thereof-does to a discipline. Lamanna and Hart (1968) in the absence of a theory opted to take a cautious stand and emphasize the lack of generaliz­ ability of their finding, because of the presumed exception, and thereby lost the important informational content of their study. Under the guidance of a theory an important gener­ alizable phenomenon is emerging from their data with the understanding that any effect studied under conditions of an unfavorable ratio between observation period and time to effect will deviate from linearity.

1.5.2  Compound 1.5.2.1  Primary Compounds Compounds show a tremendous range of inherent toxicity. Pesticides constitute only a small proportion of all indus­ trial chemicals, but even pesticides show a wide range of toxicity. For example, the oral toxicity of tetraethylpyro­ phosphate (TEPP) is approximately 588 times greater than that of a pyrethrum extract. However, it must not be sup­ posed that the difference depends on the fact that one of the compounds is synthetic and the other of plant origin, because the difference in toxicity is sometimes reversed. Nicotine, a plant product, is about 103 times more toxic than difenphos, a synthetic organic phosphorus compound. The most toxic materials known are produced by living organisms. Table 1.5 illustrates the range of toxicity produced by one or a few doses of selected pesticides and some other materials. There is some tendency for compounds of simi­ lar chemical nature to resemble one another in toxicity (structure/activity relationship). However, the resemblance is more likely to be quantitative than qualitative. Thus, the organic phosphorus compounds all produce a similar clini­ cal picture, but difenphos does so only at a dosage over 10,000 times greater than TEPP. The toxicity of each com­ pound must be judged separately. Compounds also show variation in inherent toxic­ ity when given repeatedly. Butler (1965) reported that aflatoxin, a poison elaborated in food by certain fungi, produces cancer in rats at a dosage of only 0.01 mg/day, whereas the synthetics dimethylnitrosamine and butter yel­ low require dosages of 0.75 and 9.0 mg/day, respectively, to produce the same effect. The fact that some compounds are inherently likely to produce chronic illness whereas others produce acute poi­ soning only, regardless of the duration of intake, must be reemphasized here. Again, the theory of toxicology provides an explana­ tion for this widely observed and reported phenomenon which puzzled several generations of toxicologists. What is required though, is the abandonment of semiquantitative notions of time such as acute, subacute, subchronic and

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

chronic and to replace them with quantitative measurement of time as an independent variable of toxicity. As discussed and explained in section 1.1.4 ideal conditions for study­ ing dose-time-effect relationships are either isoeffective, isotemporal or isodosic. Nature provides many examples when these conditions are met or nearly so. Cyanide is a good example for what is considered an acute poison. The reason for that is that the time to death is short and there­ fore essentially no recovery occurs between exposure and effect (death). Execution by cyanide in a chamber provided the theoretically ideal condition of steady state exposure. However, in a few instances of nearly fatal cyanide poi­ soning the individuals recovered with permanent (chronic) brain damage because of transient hypoxia. Asbestos, in contrast, is considered a chronic poison. Asbestos causes cancers (lung cancer and mesothelioma) only after chronic exposure and the dose cannot be increased to levels when acute exposure would cause chronic toxicity (cancer), because it would suffocate the animals acutely. This brief discussion illustrates that long-entreched notions can be outright harmful for the advancement of a ­ discipline because they actively prevent the development of (a) theory(ies), which is a mandatory requirement for any improvement in experimental design.

1.5.2.2  Derived Compounds Not only do compounds differ in their inherent toxicity, but they differ in the ease with which they undergo chemical change. Some pesticides may decompose during storage. Others change when their residues are exposed to ultraviolet light, plant enzymes, or soil microorganisms. Thus, one or more derivatives, in addition to the original compound, may be absorbed by humans or animals exposed in one of several ways, including exposure by eating food treated earlier by a pesticide. Of course, nearly all compounds (whether viewed as primary or derivative) are metabolized following absorp­ tion by humans or animals. No two compounds are exactly alike. Each derivative and metabolite will differ chemically and toxicologically in some degree from its precursor. There is no rule regarding the relative toxicity of com­ pounds and their nonmetabolic derivatives. Metabolism tends to render compounds more water soluble and less toxic, but there are instances when this is not the case. Peters (1952) coined the term “lethal synthesis” in 1951 for biotransformation of a compound to a significantly more toxic product, which in modern terminology is called meta­ bolic activation. Full understanding of the toxicology of each pesticide can be acquired only through recognition and study of its derivatives as well as the primary compound. Such study may reveal that the toxicity of a compound depends on a lethal synthesis. This discovery may or may not suggest the possibility of some preventive or therapeutic measure. However, in no event will discovery of the details change

55

the inherent toxicity of the primary compound. Parathion is no more or less toxic since the discovery that its toxicity depends largely on its conversion to paraoxon. The usual presence of impurities (from synthesis or decomposition) combined with metabolic conversions occurring in an organism turns practically even the purest compound into a more or less complex mixture. A chemical moiety is one of two essential elements of a toxic interaction, the other one being a subject or a popu­ lation of subjects. Toxic potency is an intrinsic property of each and all chemicals. It may be defined as the dose of a compound to cause a defined level of toxicity (ED 20, ED 50, or ED 80) at constant time (all timescales must be kept constant) for time points between the minimum lag period of an effect and the maximum life span of a spe­ cies. Because determination of relative potency at the mini­ mum lag period or maximum life span would require huge populations of experimental animals, it can be determined more conveniently at ED 20 to ED 80 under conditions of toxicokinetic-toxicodynamic steady state. As discussed for Fig. 1.5, toxic potency is not dose dependent unless the experiment is not measuring toxicity, but various ratios of toxicity/recovery. Dose-response curves are certainly always parallel for chemicals acting by the same mechanism (Stahl et al., 1992) under conditions of kinetic or dynamic steady state or as long as departure from that condition is minimal during the observation period. This is the only situation when valid structure-activity (relative potency) relation­ ships can be established. Metabolites  Biotransformation usually leads to more water-soluble (phase I) or very water-soluble (phase II) derivatives, which by definition have shorter half-lives than the parent compound. Therefore, in a chronic dosing exper­ iment the steady state concentration of the metabolite will not correspond to the percentage of metabolite formed but a fraction thereof depending on how much shorter the halflife of a metabolite is compared to the parent compound. Thus, even if a phase I metabolite has agonistic proper­ ties leading to additivity of effect, it is seldom of practical importance. In addition, phase I biotransformation products are usually rapidly converted to conjugates, which usually do not have agonistic properties with the parent compound and even shorter half-lives. These theoretical considerations are in agreement with the practical experience that bio­ transformation of chemicals most often represents detoxi­ fication (kinetic recovery). The only exception is metabolic activation or lethal synthesis as coined by Peters (1952) and cited by Hayes (1991), which is extremely well-understood compared to the frequency of its occurrence in toxicology. The questions to be asked in this case are not different in principle, except that now the metabolite rather than the parent compound is the most toxic component. If they act by the same mechanism (agonists in a broader sense than used in pharmacology) then the significance of additivity

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56

depends on the relative kinetics of the two or more com­ pounds. It should be kept in mind though, that if the halflife of a more toxic metabolite is much, much shorter than that of the parent compound, its role in, or contribution to, overall toxicity might still become insignificant due to its correspondingly lower steady state concentration. Impurities  TCDD was an impurity (a few parts per mil­ lion) in early batches of Agent Orange (2,4,5-T and 2,4-D mixture) and other chlorophenol-related products due to the synthetic process. Because of its high potency and extremely long half-life, it became virtually the only compound of toxicological concern in these products (Kimbrough, 1980). Contrary to this constellation, the pres­ ence of small quantities of chlordecone in mirex batches is of little toxicological significance, because its potency is not that much greater than that of mirex and its half-life is shorter (ATSDR, l995a). However, chlordecone is a neu­ rotoxin in its own right and given sufficient exposure will cause this type of toxicity.

Methods of Measuring Interaction  It has been pointed out that, for statistical reasons, it is possible to estimate the dosage responsible for an ED 50 more accurately than the dosage responsible for some greater or lesser effect. It is for this reason that in studying the interaction of two or more compounds, they are often administered in equal fractions of their respective ED 50 values. If two compounds are compared, the dosages should be a geometric series based on one-half, for three compounds the dosages should be a geometric series based on one-third, and so on. Thus, if the effects of two compounds are exactly additive, administra­ tion of half an ED 50 compound A and half an ED 50 com­ pound B should result in exactly one ED 50 for the mixture. This relationship may be written as



If the compounds are antagonistic by a factor of 2, the relationship may be written

1.5.3  Interaction of Compounds In a broad sense, it is probable that all compounds in the body interact, directly or otherwise. Most of the interactions are so complex, obscure, or trivial that they remain and most of them, should remain unidentified. However, some foreign chemicals have distinct interactions in the body, and in some instances the mechanisms of these ­interactions have been identified. The compounds that interact may be two or more drugs, may be two or more poisons, or may be an active ingredient(s) and one or more vehicles or other constituents of a formulation sold as a drug or pesticide. Some examples of interaction not mentioned in the fol­ lowing paragraphs may be found as part of the discussion of individual compounds or groups of compounds in other parts of this book.

1.5.3.1  Kinds of Interaction The effects of different foreign chemicals may (a) mutually interfere with one another, (b) be simply additive, or (c) potentiate one another. The essentially additive relationship is the most common. Both exceptional conditions-mutual antagonism and potentiation-may be of practical and theo­ retical importance. From a practical standpoint, interfer­ ence (antagonism) between two compounds may cancel the benefit or counteract the injury expected from one of them. Potentiation may increase benefit or harm depend­ ing on circumstances. From a theoretical standpoint, study of either antagonism or potentiation often leads to a ­better understanding of the mechanisms of action of the com­ pounds involved.

1 ( ED50 A  ED50 B ) ⋅ 1.0  1ED50 M 2

(1ED50 A  1ED50 B ) ⋅ 0.5  1ED50 M

If the effects of the compounds potentiate one another by a factor of 4, the relationship may be written



 1  ED50 A  1 ED50 B  ⋅ 4.0  1ED50 M   8 8

Written in this way the multiplicand indicates the ratio between the observed and the expected ED 50 for the mixture and expressed the degree and kind of interaction. Thus, 1.0 indicates an exactly additive relationship, pro­ gressively smaller fractions indicate progressively greater antagonism, and numbers progressively greater than 1.0 indicate progressively greater potentiation. Actually, the error of measurement is such that fractions or numbers dif­ fering from 1.0 by no more than a factor of 2 or 3 cannot be distinguished from the simple additive relationship. In some instances, it is desirable to study the interac­ tion of two compounds at many dosage ratios. The results of such tests may be recorded in a diagram such as that shown in Fig 1.17. This kind of diagram was introduced in 1926 through a paper devoted to theoretical and math­ ematical considerations (Loewe and Muischnek, 1926) and another dealing with the antagonism between barbital and aminopyrine (Kaer and Loewe, 1926). Loewe and Muischnek (1926) introduced the term “iso­ bole” (from the Greek isos, equal, and bolos, a blow or stroke) to designate a line passing through points of equal action or injury, for example, a series of ED 50 values result­ ing from administering two compounds in different ratios. In Fig. 1.17, the dotted line indicates all possible compari­ sons at equal ratios of the two ED 50 values. The three time points on the diagram indicate the same ­ relationships of dosage as those presented by the three equations in the

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

57

100

An t

ag

on

is m

80

tio

n

60

en

tia

40

Po t

Percentage of ED 50 of compound B

120

20

20 40 60 80 100 120 Percentage of ED 50 of compound A Figure 1.17  Isoboles of ED 50 values of compounds A and B illustrating additive, antagonistic, and potentiative interactions. See text for further explanation.

28

ED 50 sulfadiazine (mg/kg)

24 20 16 12 8 4

0

4

8

12

16

20

24

28

32

36

ED 50 pyrimethamine (µg/kg)

Figure 1.18  ED 50s (dosages reducing parasitemia to 50% of the in parasitemia of untreated controls) of pyrimethamine and sulfadiazine, adminis­ tered both singly and together in various proportions to chicks infected malaria. Each ED 50 was determined graphically from a dosage–response curve. Redrawn from Rollo (1955), by permission of the British Journal of Pharmacology.

preceding paragraph. The solid straight line is the isobole of exactly additive action at all dosage ratios. All lines (includ­ ing one shown) lying to the right and above the isobole represent some degree of potentiation. Real curves are not always symmetrical.

A number of other theoretical relationships or special cases in addition to additive, antagonistic, and potentiative have been mentioned, but whether they exist in nature or can be meaningfully distinguished by the type of diagram shown in Fig. 1.17 is not clear. Examples include a combination

58

of antagonism and synergism between the same pair of compounds at different dosage ratios, sensitization, and desensitization. The axes of the diagram need not be measured in percent­ age, but may indicate dosage directly as shown in Fig. 1.18, which records the toxic (therapeutic) effect of pyrimethamine and sulfadiazine, administered both singly and together in various proportions, to combat malaria organisms infecting chicks. Systematic Study of Interaction  Only a few studies have been made of the possible interactions of compounds in such a way that the results can be compared meaningfully. One such study was that of Keplinger and Deichmann (1967). It involved over 100 combinations of eight chlo­ rinated hydrocarbon insecticides, six organic phospho­ rous insecticides, and one carbamate insecticide in sets of two and three compounds in rats and mice. The results were expressed as the quotient of the ratio of the expected to observed LD 50 values. The largest quotient obtained was only 2.26, indicating a very small or even question­ able potentiation between chlordane and methoxychlor in mice. The smallest quotient obtained was 0.36, indicating a minor degree of antagonism between aldrin and trithion in rats. The data for combinations of three compounds did not reveal any effects of toxicity that could not have been predicted from the combination of two compounds. In a 2-year rat-feeding study using a combination of six pesticides (DDT, aldrin, pyrethrin, piperonyl butoxide, malathion, and 2,4-D) and eight food additives at use level concentrations or higher, significant alteration of toxicity in comparison with the toxicity of individual substances was not found (Fitzhugh, 1966). For industrial chemicals as well as for pesticides, the most common joint action is a simple additive one (Smyth et al., 1969). Antagonism  Although some instances of interference are encountered in general surveys of interactions, the exam­ ples usually show interference that is both small in magni­ tude and unexplained. A rapid change in the degree of antagonism may be of more clinical significance than antagonism per se. For example, Cucinell et al. (1966) reported a fatal hemorrhage in a patient who had received chloral hydrate and bishy­ droxycoumarin in combination without ill effect. However, when medication with chloral hydrate was stopped but bishydroxycoumarin was continued, the prothrombin time increased and hemorrhage occurred. It was later shown that chloral hydrate stimulates the metabolism of this anticoagu­ lant. The danger lies in too rapid withdrawal of the inducer without an appropriate reduction in the dose of the antico­ agulant. The same danger does not exist in connection with inducers that are stored to a significant degree in the tissues because, even if their administration is discontinued,

Hayes’ Handbook of Pesticide Toxicology

their action decreases very gradually because of their slow elimination from the tissue. Potentiation  A few examples of clinically important potentiation are known to involve pesticides. Many chlori­ nated hydrocarbon solvents and fumigants, notably carbon tetrachloride, are much more likely to injure the liver if alcohol is consumed at the same time. The hepatotoxicity of many haloalkanes is potentiated by many compounds that induce microsomal enzymes of the liver, by exogenous ketones, including chlordecone, or by metabolic ketosis (Hewitt et al., 1980). By a totally different mechanism, the dithiocarbamate fungicides, which are closely related to disulfiram, interfere with the metabolism of alcohol so that alcohol becomes more toxic. True potentiation is a comparatively rare phenomenon except in connection with certain organic phosphorus insec­ ticides and at least some classes of teratogens. The reason for the interaction of organic phosphorus compounds is that many of them inhibit aliesterases responsible for the ­efficient detoxification of some other members of the same class (Su et al., 1971). This is the mechanism that explains the potentiation of the toxicity of malathion by ethyl p-nitro­ phenyl thionobenzenephosphonate (EPN) when the dos­ age of both is substantial (Murphy and DuBois, 1957). However, if the dosage of the two compounds is sufficiently small, there is enough enzyme to detoxify both of them, and the phenomenon of potentiation is not manifest. Rider et al. (1959) have shown that people can tolerate 3 mg/day of EPN plus 16 mg/day of malathion or 6 mg/day of EPN plus 8 mg/day of malathion for prolonged periods without significant depression of red cell or plasma cholinester­ ase. The combination of 6 mg/day of EPN plus 16 mg/day of malathion (Rider et al., 1959) or 16 and 5 mg/person/ day, respectively (Moeller and Rider, 1962a), did produce asymptomatic depression of both enzymes, but the effect was only additive. No potentiation was noted. The high­ est dosage of malathion alone tolerated without even slight inhibition of cholinesterase is 16 mg/person/day (Moeller and Rider, 1962a), whereas that for EPN is 6 mg/person/day (Rider et al., 1959). Thus, potentiation among this class of compounds may be important for overexposed workers but not for people who ingest residues on foods. Other mechanisms of interaction are outlined in Section 1.5.3.2. In a study of the interaction of six recognized teratogens, it was found that all pairs showed appreciable potentiation of teratogenic action provided the dosage of each was above a level producing at least a 1 % effect. In several instances, potentiation occurred even when one or more materials were given at sub-threshold dosage. No such consistent pattern of interaction was observed regard­ ing intrauterine death (Wilson, 1964). It seems likely that the mechanism of cocarcinogenesis will not be explained until neoplasia itself is explained. However, in what may have been the only quantitative

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

study of this kind of potentiation, a dosage-response relationship was found not only for 9,10-dimethyl- 1,2­benzanthracene but also for this compound in combination with the cocarcinogen croton oil (Graffi, 1953). Whether potentiation will be of any practical impor­ tance depends on its degree and on the chance a person or animal may have simultaneous exposures to adequate amounts of two potentiating compounds. Potentiation as high as four- or five-fold, such as that seen with some organic phosphorus insecticides, is of limited toxicologi­ cal importance. The chance that a person or animal will encounter both members of a potentiating pair is smaller than the chance of encountering either one separately. This is especially true because, to be effective, the two com­ pounds must be absorbed at about the same time and at dosages not very different from those that would be dan­ gerous if only a single compound were absorbed. Traces are not effective. Potentiation may be of critical importance in isolated instances, but it is virtually impossible to predict it, partly because there may be no apparent-and therefore ­predictivepharmacological relationship between the two compounds involved and also because the mechanism of their inter­ action (Section 1.5.3.2) may not be known initially. An exception involves the organic phosphorus insecticides. DuBois et al. (1968) developed a quantitative procedure for measuring the potency of these compounds to inhibit aliesterases and amidases that are critical to their detoxi­ fication. DuBois (1972) has suggested that the use of this procedure constitutes a practical method of determining the dietary levels that might potentiate the toxicity of phar­ macologically active compounds normally detoxified by esterases. A factor greater than 100 was found for potentiation between malathion and triorthocresyl phosphate (TCP) (Murphy et al., 1959). Although TCP is an organic phos­ phorus compound, it is not a pesticide. Potentiating com­ pounds need not belong to the same chemical class. An example is the potentiation of the toxicity of parathion by chlorophenothiazine (Gaines, 1962). Furthermore, the chance of encountering two compounds at about the same time is not always random. Striking exceptions are com­ pounds used in related procedures, including drugs taken concurrently or used to treat intoxication. The reason for emphasizing drugs is that many are taken at dosages that equal or exceed the daily dosages of pesticides absorbed by the most exposed workers. The exact opposite of potentiation is expected when exposure to a toxicant and its antidote are associated either for prophylaxis or treat­ ment. However, less thoroughly studied combinations of toxicants and drugs might prove to be potentiating, par­ ticularly if the drug is taken once the toxicant has been absorbed in sufficient dosage to produce illness. In this situation, even a moderate degree of potentiation might prove critical.

59

It is important that the clinician is aware of the possi­ bility of interaction and that appropriate studies be made in all cases in which poisoning appears to have occurred but in which the degree of known exposure seems inadequate to account for the observed effect.

1.5.3.2  Mechanisms of Interaction Compounds interact in the body by a wide range of mecha­ nisms, including chelation, alteration of ionization, altera­ tion of protein binding, and the inhibition, reactivation, or induction of enzymes. Original access to the body may be altered by some of these same mechanisms or by solvents, ion-exchange resins or absorptive colloids, or a change of the intestinal flora. The final pharmacological or toxico­ logical effect of one or more interactions usually cannot be predicted except by careful study of a pair of compounds. The mechanisms involved in interaction of compounds are described at greater length in other sections. The possible complexity of interactions must be empha­ sized. For example, calcium disodium EDTA is ­useful for removing lead from the body, but treatment that is too intense or prolonged or that employs certain other chelat­ ing agents can cause injury by disturbing the distribution of essential trace metals in the tissues. In their net effects, charcoal and ion-exchange resins are similar to chelating agents. The discussion of protein binding includes an illustra­ tion of competition for binding sites as the basis for the interaction of two compounds. The action of several pesticides depends on the inhibition of enzymes. The success of several antidotes depends on their ability to reactivate these inhibited enzymes. Thus eth­ ylenediaminotetracetic acid (EDTA), British anti-Lewisite (BAL), and other chelating agents may restore enzymes blocked by heavy metals. Oximes such as pralidoxime chloride (2-PAM) may restore enzymes blocked by organic phosphorus compounds. Combined use of nitrites and sodium thiosulfate releases cytochrome oxidase blocked by cyanide. Enzyme Induction  When the same substrate and enzyme are involved, inhibition and induction have opposite phar­ macological effects. Thus, inhibition of liver S-desulfurase by SKF 525-A or by feeding a protein-free diet antagonizes the action of azinphosmethyl. Conversely, induction of the same enzyme by 3-methylcholanthiene or 3, 4-­benzpyrene potentiates the action of the insecticide (Murphy and DuBois, 1958). Under different circumstances (especially the involvement of an enzyme of opposite pharmacologi­ cal action) the effect of inhibition and induction may be reversed. For example, inhibition of liver aliesterase by EPN, TCP, or a number of other compounds potentiates the action of malathion (Murphy et al., 1959; Murphy and Cheever, 1968).

60

Perhaps the decision tree will help to sort out some of the most difficult problems of toxicology and that is how to deal with mixtures of chemicals. Part of the problem might be that in spite of widespread awareness of kinetic interac­ tions there is a nearly complete absence of kinetic consider­ ations when interactions between chemicals are viewed, as exemplified by the preceding subsections or by other com­ prehensive reviews (Pöch, 1993). No attempt will be made here to address the issues of potentiation, synergism, and other complex interactions, which may be of phenomeno­ logical origin and therefore of limited theoretical ­interest. Keeping these conceptual restrictions in mind, there are two possibilities: two or more chemicals either do interact at an experimentally measurable level or else are consid­ ered to act independently. Independent action allows for a simplified safety assessment in that determining the safety of the most toxic component (cthreshold  k/tlifespan) will automatically provide protection for all other constituents of the mixture, because there are no such dose responses that would make chemicals more potent at low than at high doses (Figure 1.5). When two or more chemicals do inter­ act, they can act in concert (agonists) or against each other (antagonists). The interaction can have predominantely kinetic or mainly dynamic elements, with the possibility of interactions between the two, which probably gave rise to the notion of potentiation, synergism, etc. Kinetics  One chemical can potentially affect any step in the disposition of another chemical, leading to kinetic ago­ nism or antagonism. For example, it has been demonstrated clinically that administration of a penicillin-­sulfonamide mixture to premature infants resulted in kernicterus, because the sulfonamide displaced bilirubin from its albu­ min binding site, increasing the free fraction of bilirubin in plasma, making it thereby available for diffusion into the brain (Silverman et al., 1956). Kinetic antagonism leading to therapeutic agonism was demonstrated in the early days of penicillin when penicillin was administered together with probenecid to block its active tubular secre­ tion. Of toxicological relevance is also the fact that some organic acids compete with uric acid for the same organic acid transporter in the kidney and thereby can precipi­ tate an acute attack of gout. Still another example is the increased nonbiliary intestinal excretion of some lipophilic chemicals across the gastrointestinal wall by oral admin­ istration of mineral oil or the trapping of biliary metabo­ lites by cationic or anionic resins mixed with the feed (Rozman, 1986). Dietary constituents, vehicles, etc. often significantly alter absorption by either enhancing (kinetic agonism) or reducing (kinetic antagonism) it. These examples represent a tiny fraction of what is known about kinetic interactions between chemicals in mixtures. Yet we (Rozman and Doull) are not aware of any attempt to con­ ceptualize the role of kinetics in the toxicity of mixtures.

Hayes’ Handbook of Pesticide Toxicology

Dynamics  The effect itself is always of primary inter­ est, even when a kinetic process(es) represents the rate­determining (-limiting) step(s) (see Section 1.1.1 for the fundamental equation of toxicology). Nevertheless, the nearly complete absence of kinetic considerations when conducting tier testing (NRC, 1988) or constructing iso­ bolograms (Pöch, 1993) is lamentable and is probably part of the reason neither one is working particularly well in any other than specific situations. Classical agonistic and antagonistic interactions of binary and some ternary mix­ tures of drugs and other chemicals have been described so many times that even a superficial discussion of this topic appears unnecessary. The critical question though remains and concerns the mechanism of action. Chemicals exerting their toxicity by independent mechanisms can be dealt with by identifying the most toxic component and establishing a safety assessment for this compound, which will provide safety for all other, less potent constituents of the mixture. Chemicals acting by the same mechanism will display additivity in their effect. Antagonistic effects occur seldom because mixtures seldom contain comparable relative con­ centrations of an agonist and an antagonist. In addition, if the kinetics of the agonist and antagonist are very different, any potential interaction may turn out to be insignificant. Supraadditivity usually and perhaps always amounts to a lack of understanding of the interaction. For example, the well-known 3-10-fold synergism in organophosphate (OP) poisonings is almost certainly due to two variables, one being the ratio of the half-lives of the two OPs, the other being related to the nonspecific detoxification pool (plasma carboxyesterases and other high affinity proteins in tissues). If both OPs are administered as single compounds, a large percentage of the dose of each will be detoxified by plasma carboxyesterases and only a fraction of the dose will reach the primary targets of acute toxicity (central nervous sys­ tem, lungs, diaphragm). However, if one OP is admin­ istered before the other, then (depending on timing) the second OP will encounter varying degrees of occupation of the detoxification sites in plasma and therefore a larger portion of its dose will be available to exert toxicity at the target sites. It is our conviction that, although it may not be possible to explain potentiation and synergism on grounds of dose responses alone, but we may very well be able to do so in terms of both dose and time as variables of the interaction.

1.5.3.3  Interactions that May Influence Laboratory Tests Most commercial products that are used as pesticides are marketed as formulations and thus contain vehicles and other ingredients that give the pesticide the desired ­properties for its intended use. These materials that are added to the active ingredients to provide the proper physical characteristics

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

cannot be assumed to be inert; these materials have toxico­ logical properties of their own that are, in some cases, of greater importance than the toxicological properties of the active ingredient itself. In some of those cases, the toxicity of the total formulation reflects that of the active ingredi­ ent only because it is there in larger quantity than the other ingredients. Similarly, in toxicological studies, vehicles and other materials that may be added to a test agent have some influence on its toxicity. Except for studies of toxicity to the eye, most toxicological studies do not use the application of chemicals in their pure form. Inhalation toxicology studies may be done on a pure chemical, but often are conducted using technical products or commercial products, which have stabilizers or other ingredients that may or may not influence the toxicity of the chemical under test. In toxico­ logical studies by other routes of administration, vehicles of one kind or another are routinely used to provide good mixing in diets or drinking water or to provide consistency of dose volume and ease of measurement of doses admin­ istered. The actual vehicle used varies widely by type of study, route of administration, nature of the chemical to be administered, as well as geographic region of the world. In the United States, the most commonly used vehicles are water and corn oil. Because most drugs and other organic chemicals in commercial use are not very soluble in water, an oil-based vehicle is widely used commercially and in tox­ icology laboratories. Effect of Formulation  The toxicity of a compound may be modified by differences in formulation. Solvents are espe­ cially important in this connection, but wetting agents and other ancillary compounds may be involved. When these chemicals promote or retard the toxicity of a pesticide, it is usually through promotion or retardation of absorption. The facilitating action may involve injury to a barrier, especially in the skin. Increase in absorption may also involve a sol­ vent that, by its own ready absorption, enhances absorption of the toxicant. Importance of Environmental Chemicals  The source of a compound that influences the toxicological or pharma­ cological action of a recognized compound is not always obvious. A striking example is the alteration of drug metabolism in rats and mice by cedarwood bedding in their cages (Ferguson, 1966; Vesell, 1967; Wade et al., 1968). Another example is the change in reaction to molybdenum caused by traces of zinc derived from galvanized cages (Section 1.5.11.4). An example of a toxic interaction from human experience that was at first obscure is that of asbes­ tos dust and cigarette smoke. The identity of some other interactants is obvious. The concentration of ammonia fumes in the air of animal rooms from bedding soiled with urine has occasionally been a source of complaint by personnel working in the animal

61

rooms. It is also now recognized as a possible complicating factor in the interpretation of animal studies, particularly when there might be respiratory lesions. Broderson et al. (1976) evaluated the effects of ammonia at concentrations of 25–250 ppm in the air of animal rooms on the character­ istics of murine respiratory mycoplasmosis in Sherman and Fischer rats. The prevalence of pneumonia, but not of other respiratory lesions of murine respiratory mycoplasmosis, showed a strong tendency to increase directly with envi­ ronmental ammonia concentrations. Exposure to ammo­ nia of rats that had not been infected with the mycoplasma organism caused anatomic lesions that were unlike those of mycoplasmosis and were limited to the nasal passages. The authors concluded that environmental ammonia at concen­ trations commonly encountered in cage environments for rats played an important role in the pathogenesis of murine respiratory mycoplasmosis. Some information on detectable concentrations of bac­ terial toxins, heavy metals, solvents, pesticides, and other environmental contaminants in laboratory feed and drink­ ing water is available (Newell, 1980; Rao and Knapka, 1987; Rao, 1986; Williams, 1984).

1.5.4  Schedule of Dosage It is common knowledge among toxicologists that the schedule of dosage may have an important influence on the quantitative results. Usually anything that permits greater detoxification or excretion of a toxin tends to reduce the injury it produces. An oral dose given on an empty stomach is absorbed over a briefer period than the same dose admin­ istered when the stomach is at least partly full. Ingestion of a certain daily dosage mixed in the diet often is less injuri­ ous than the same dosage of the same compound adminis­ tered daily by stomach tube. The compound reaches a lower maximal concentration in blood and other tissues when the same dosage is distributed throughout the day rather than concentrated in a brief period of time. The microsomal enzymes, excretion, and other defenses may be able to cope indefinitely with a low concentration of a compound but may not be capable of handling peak levels. Similar reasoning applies to schedules that permit rest periods as compared with those that do not. Truly continuous exposure is usually more damaging than intermittent expo­ sure at the same daily rate. An example may be cited for lead (Kehoe, 1961). It must be noted, however, that the distinc­ tion between continuous and intermittent exposure is blurred somewhat for a compound that is stored, such as lead. An even more dramatic example involves carbon tetra­ chloride studied in connection with the possible continuous exposure of people in submarine vessels or stations. It was found that intermittent exposure (8 h/day, 5 days/week) to carbon tetrachloride at a concentration of 515 mg/m3

62

killed a small proportion of experimental animals and caused injury, especially to the liver, of many of those that survived for 6 weeks. About the same degree of injury was produced by continuous exposure at a concentration of only 61 mg/m3, and this occurred within about the first 6 weeks (Prendergast et al., 1967). Thus, under these con­ ditions, continuous exposure was about eight times more dangerous based on concentration and twice as dangerous based on total dose than intermittent exposure similar in schedule to much occupational exposure. The effects of different schedules of dosing may differ qualitatively as well as quantitatively, sometimes in such a way that intermittent exposure to a concentration too high to tolerate continuously leads to a greater variety of pathol­ ogy than is seen under any other condition. Thus, Landry et al. (1985) reported that mice exposed to methyl chloride at 2400 ppm for only 5.5 h/day showed renal pathology, intravascular hemolysis, and hematopoietic effects in addi­ tion to the cerebellar granular cell degeneration and con­ sequent neuromuscular dysfunction seen in mice exposed to lower concentration on the same schedule or in those exposed to any of a range of concentrations for 22 h/day. The cited work by Prendergast and his colleagues and earlier related work by the same and other groups of inves­ tigators offer some indication that a considerably smaller dosage of each of a number of compounds is required to cause injury if exposure is continuous rather than intermit­ tent. Unfortunately, many of the investigations are reported in such a way that a meaningful comparison is impossible. Because many people are exposed to air pollution, some of them continuously, it seems tragic not to compare the effects of continuous and intermittent exposure at equal intervals of time after initial exposure. This is particularly true because the equipment and procedure for continuous exposure are specialized and costly. As long as tests are to be done, little difficulty or expense would be added by gathering and presenting comparable data. Certainly the number of persons now exposed continuously to ordinary air pollution is vast compared with the number who will enter the closed atmospheres of spacecraft or submarine vessels or stations in the foreseeable future. The general rule that rest periods and avoidance of peak blood levels tend to be protective usually applies most to compounds that are easily detoxified and excreted, and least to compounds against which the defenses of the body are inherently poor, with the result that the compounds or their effects are relatively cumulative. The cumulation may occur over relatively long periods as, for example, with lead, or over short periods as, for example, with carbon monox­ ide. In the former case, the cumulation frequently involves months. In the latter case, the cumulative effect may involve hours or days, but is not prominent in connection with longer ­periods. All of us inhale carbon monoxide, and those of us who smoke tobacco inhale more than nonsmokers (Hanson and Hastings, 1933). Some garage workers encounter a level

Hayes’ Handbook of Pesticide Toxicology

of exposure that is marginal with respect to injury. Higher levels of exposure involve progressively more hazard with the result that in some countries carbon monoxide kills more people than any other single compound. In some instances (Saffiotti and Shubik, 1956; Taylor and Nettesheim, 1975; Waud et al., 1958) repeated small doses produce greater effects than a smaller number of larger doses even though the total dosage resulting from the larger number of applications is the same or less. This relationship would appear to violate a dosage-response relationship. The explanation may involve a failure of one or a few doses to reach the target tissue. Prolonged action may be required when there is an inherent delay between the initial dose and first observed effect regard­ less of whether this effect follows one or more doses. The apparently inverted dosage response may involve a purely pharmacological effect such as the depletion of tissue nor­ epinephrine by reserpine (Waud et al., 1958) or it may involve toxic effects as discussed in Section 1.3.3.2. The modifying effect of schedule and a number of other factors must be taken into account in any considerations of dosage. Tests to establish safe levels should involve the dos­ age schedule, route, and other conditions people are expected to encounter. When these modifying factors are taken into account, the paramount importance of dosage becomes even more evident. This section illustrates best how well outstanding toxi­ cologists understood the factors influencing toxicity. Yet, it was not possible to take this discipline to the next level without recognizing time as a independent quantitative and quantifiable variable of toxicity. The example of carbon tetrachloride toxicity in rats used by Hayes is a good case to illustrate the power of examining both dose and time as variables of toxicity. The half-life of carbon tetrachloride is about 7 h (ATSDR, 2005). Exposure to 515 mg/m3 for 8 h means that rats were continuously exposed to this compound for 24 h/6 weeks above about 130 mg/m3, which is clearly above the tox­ icity threshold because 61 mg/m3 for 8 h/6 weeks also caused moderate liver pathology. However, the later expo­ sure dropped to about 12 ppm after every day’s 8 h expo­ sure, which was just slightly above the 7 h hepatic toxicity threshold (8 ppm) (see ATSDR, 2005) and below the 6 h/day for 4 days LOAEL of 50 ppm. Therefore, it is very obvious that at the higher dose rate (515 mg/m3) rats were afforded no time for recovery between exposure episodes, whereas at the lower dose rate they had plenty of time for recovery. Haye’s other examples are similarly easy to explain using the theory of toxicology.

1.5.5  Duration of Dosage Weil and McCollister (1963) investigated the degree of toxicity revealed by short-term and long-term tests in rats.

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

For 22 compounds, the ratios of the dosages producing the minimal effect observed in short-term and in 2-year feed­ ing tests varied from 0.5 to 20.0 and averaged 2.9. Ratios greater than 1.0 indicated the degree of apparent increase in toxicity associated with long-term testing. Ratios less then 1.0 may have indicated adaptation, experimental variation, or both. It was possible to compare the maximal dosages of 33 compounds producing no effect; these ratios ranged from 0.5 to 12.0 and averaged 2.3. Some methods for measuring the effects of duration of dosage are discussed in Sections 1.3.1.2, 1.3.1.3, 1.3.2, and 1.3.3.2. Similar studies were continued by Weil et al. (1969) and expressed in somewhat different terms. The comparisons involved 20 compounds, including 11 pesticides. The LD 50 values were determined and the compounds were fed to rats for 7 and 90 days, respectively. The results are com­ pared with those in 2-year studies done earlier. The LD 50 values offered a poor indication of the results of repeated dosing. However, the results of long-term exposure could be predicted in an efficient way from the results of exposure lasting only 7 days. Using subscripts to indicate the num­ ber of days of exposure, it was found that the relationships for predicting the lowest dosage large enough to produce a minimal effect (MiE) were those shown in Table 1.6. Littlefield and Gaylor (1985) showed that under condi­ tions of the study, daily dosage rate (mg/kg/day) seemed to be more important than duration of dosing in increas­ ing the prevalence of liver and bladder tumors in mice fed 2-AAF when the total dosage was the same (Rozman and Doull, 2001a).

1.5.6  Route of Exposure The route by which a compound is absorbed helps to deter­ mine not only the ease of absorption, but also, in some instances, the ease of metabolism. Compounds are ­usually more toxic by the oral then by the dermal route. This was true of 64 of 67 compounds studied by Gaines (1960, 1969) and analyzed in this regard by Hayes (1967a). However, there were three exceptions, that is, three compounds more toxic by the dermal route. Considering all 67 compounds, Table 1.6  Ratios for Predicting the Results of LongTerm Feeding from the Results of Short-Term Feedinga Value

Ratios for predicting result of 90-day feeding study

2-year feeding study

Median value

MiE 7/3.0

MiE 90/1.8 or MiE 7/5.4

95th percentile

MiE 7/6.2

MiE 90/5.7 or MiE 7/35.3

a

Modified from Weil et al. (1969), by permission of Academic Press.

63

the factor of difference by which oral toxicity exceeded dermal toxicity raged from 0.2 to 21 and averaged 4.2. The lesser toxicity of one of the compounds (isolan) by the oral route was markedly influenced by metabolism. Five of six rats survived infusion of isolan into an intestinal vein for an hour at a rate that led to death within 18–35 minutes in six comparable animals infused via the femoral vein (Gaines et al., 1966). Thus, a single pass through the liver is sufficient to make the difference between life and death as a result of exposure to isolan. This phenomenon helps to explain the high dermal and low oral toxicity of the compound. The high dermal toxicity of monochloracetic acid was due to ­irritationrelated rapid absorption and the much lower oral toxicity occurred because of delayed stomach emptying, also due to local irritation (Saghir and Rozman, 2003). Gaines (1969) found that about one-third of the pesti­ cides he tested had such a low or variable dermal toxicity that no LD 50 could be determined. Thus, the true average difference between oral and dermal toxicity is greater than that calculated for compounds for which definite dermal as well as oral LD 50 values can be measured. Furthermore, relatively low dermal toxicity may be characteristic of an even higher proportion of compounds generally than is true of pesticides. For practical reasons, respiratory toxicity usually is studied and reported in terms of concentration of chemical and duration of exposure. Values obtained in this way can­ not be converted easily to dosage in terms of body weight, and direct comparison with toxicity by other routes is not usually possible. Some notion of the respiratory toxicity of a compound may be inferred from its intravenous tox­ icity (DuBois and Geiling, 1959). However, the method is of limited value partly because gases and aerosols are absorbed by the respiratory tract to different degrees depending on the compound and the particle size. The route of administration may have a clear effect on the delayed neurotoxicity induced by certain ­organophosphorous esters. As summarized by Francis (1983), TCP caused delayed neurotoxicity in rhesus and squirrel monkeys and dogs when given by subcutaneous administration but not by the oral route. Mipafox caused a positive response when given to rats subcutaneously but caused only an equivocal response when given in the diet. These differences may have been related to differences in metabolism involving a single pass through the liver. The differential in response from two different modes of oral administration, diet versus gavage, was evaluated by Weil et al. (1973). Rats in a reproduction study and guinea pigs in a teratology study were given carbaryl by one of the modes of oral exposure. Maximal dosage lev­ els in the reproduction study were 200 mg/kg/day by diet or 100 mg/kg/day by gavage. Maximal dosage levels in the teratology study were 300 mg/kg by diet or 200 mg/kg by gavage. The maximal gavage groups had severe maternal toxicity in contrast to little or no effect in the diet groups

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64

that received higher dosages. Thus, differences in the mode of oral administration (gavage versus diet), which is essen­ tially a difference in schedule of dosage, can be as impor­ tant as differences between routes of exposure. Another clear example of an effect of route of adminis­ tration and dosage schedule on toxicity is a study reported by Taylor and Nettesheim (1975) regarding the evaluation of the carcinogenicity of nitrosoheptamethyleneimine. F344 and Sprague-Dawley rats were given this chemical by gavage or subcutaneous injection; cumulative dosages ranged from 5.5 to 1200 mg/kg and the dosage schedule ranged from 40 serial administrations to one single injec­ tion. Oral administration was more effective in producing tumors than was subcutaneous injection at approximately the same total dosage, and administration of multiple small doses was more effective than a single large dose when the total dosage was constant. All three preceding subsections relate to time as a variable of toxicity. As such, they belong together with Section 1.5.1 (dose) dealing with quantitative variables of toxicity. Schedule and duration of dosing has been com­ bined and expressed as exposure frequency (Section 1.1.1) whereas the route of exposure is related to the timescale of absorption. These topics have been discussed in great detail in Section 1.1.1 and here it should suffice to illustrate this with still another example how the theory of toxicology together with the decision tree helps explain experiments that were misinterpreted by both toxicologists and non­ toxicologists. Littlefield and Gaylor (1985)’s interpretation of another part of the ED 01 study is a good example of nontoxicologists addressing issues they do not understand. This particular paper claims that the daily dose (dose rate) is more important in the bladder carcinogenicity of 2-AAF than is the total dose (dose), because mice receiving a high dose rate (150 ppm) of the compound for 9 months had a higher cancer rate at 18 and 24 months than those given a lower dose rate (60 ppm) for 24 months. First, the dose (total dose) administered to the mice was comparable in the two groups of animals as demonstrated by the AUCs of exposure of 150 ppm  9 months  1350 ppm    months for a high dose group or 60 ppm  24 months  1440 ppm   months for a low dose group of animals. Yet mice in the lower dose rate group had 1% bladder cancer each at 18 and 24 months, whereas mice in the higher dose rate group after 6 and 12 months recovery still had much higher cancer rates (6 and 18%, respectively). This finding is entirely consistent with the theory of toxicology. Polycycic aromatic hydrocarbons have short kinetic half-lives (about 1 day or less), but long dynamic (recovery) half-lives which dominate (rate-determining step) their action. Thus, the dynamic AUC kept growing for the group for which dos­ ing was stopped at 9 months and as a consequence blad­ der cancer incidence increased from 6% to 18% during the period lasting for 18–24 months or for any of the time 



periods after cessation of dosing (Littlefield and Gaylor, 1985). This increase was entirely consistent without regard to whether dosing was stopped at 9, 12, 15, or 18 months into dosing. Thus, it is very clear that the higher dose rate generated a much larger dynamic AUC than the lower dose rate. Deducting the threshold AUC from both the high and low dose dynamic AUCs indicates that 150 ppm 2-AAF administered for 9 months is 5–10 times more effective after 18 and 24 months in terms of dynamic AUC than is 60 ppm 2-AAF administered continuously for these periods of time. This example is a further illustration of how the theory of toxicology allows for a biologically plausible interpretation of results that, in the past, usually were submerged in a maze of biologically implausible but formalistically correct statis­ tical gibberish.

1.5.7  Species and Strain Differences This section deals almost exclusively with dynamic aspects of species and strain differences with only peripheral ref­ erence to species differences that are due to differential kinetics, although knowledge of the latter is very wide­ spread. This discrepancy illustrates that, unless knowledge is conceptualized in the framework of a theory, its applica­ tion remains haphazard.

1.5.7.1  Species Differences Due to Dynamics It is well recognized that species differences impose con­ siderable limitations on our ability to predict the toxicity of a compound from one species to another. The dismal state that has resulted from predicting that a compound found to be carcinogenic in the rat will also be carcinogenic in the mouse and vice versa is a case in point (DiCarlo, 1984; DiCarlo and Fung, 1984). Therefore, many scientists and even more nonscientists question the value of whole-­animal studies for the protection of the public from potential adverse health effects of drugs and other chemicals. Most of the scientific arguments in support of moving away from whole-animal studies are firmly rooted in the pre­ vailing reductionist thinking of the 20th century. The quintes­ sence of these arguments is an irrational hope that molecular events at the level of DNA and RNA and/or at high affinity protein binding sites will eventually explain everything that goes awry in an organism as a result of a toxic insult. In sharp contrast and with barely audible voice at pres­ ent are a few antireductionists, who despair at the complex­ ity of a mammalian organism. In their view, the multitude of causes for species differences in toxicology precludes the possibility of resolving these issues. However, whole animals still represent a more valid biological system for comparison and prediction because of some qualitative similarities between species.

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

Both sides, of course, have some valid points, but they miss many more. The overwhelming importance of genetics in the inherent capability of a cell to respond to a toxic insult has been demonstrated amply. However, the cell is not a pile of molecules and an organism is not a random cell culture, rather, cells are organized in tissues in a hierarchial fash­ ion with implicit rank order of importance for the organism. Moreover, there is a flow of information not only between adjacent cells but also between tissues. Therefore, the ques­ tion of how a particular species will respond to a toxic insult depends not only on the interaction between a xenobiotic and a subcellular element, but also on the hierarchical sta­ tus of the target tissue and on the possible disturbance of information flow between tissues. Having said this much, it is clear that a resolution of species differences will have to involve both analytical and synthetic thinking. Moreover, although understanding of species differences is important, it is so only to the extent of defining similarity (species- or strain-reactivity) between species to allow meaningful inter­ pretation of results from one species to another. Differences between Parasites and Hosts  There is a ten­ dency to ignore, as objects of scientific interest and wonder, the differences in the susceptibility of pests and of organ­ isms we hope to protect. This attitude may be justified if the difference is based more on difference in exposure than on inherent susceptibility. The difference cannot be ignored when it involves “systemics,” that is, pesticides used as drugs to combat parasites on or in their host. Examples of systemics for mammals include crufomate, trichlorfon (met­ rifonate), dichlorvos, ronnel, diphacinone (diphenadione), and coumaphos. Some of these compounds were originally developed to destroy botflies that pass their larval stages in the tissues of cattle, the mucous membranes or nasal sinuses of sheep, or the stomach and anterior small intestines of horses. Ronnel is used as a systemic treatment for fleas and the action of dichlorvos on fleas may be partially systemic. During the early studies, it was found that control extended to some but not all species of nematodes, including some in the tissues rather than the intestinal lumen. Trichiorfon has been used to treat helminthiasis, including ankylosomiasis, ascariasis, trichuriasis, and creeping eruption in humans (Cerf et al., 1962). The expected pharmacological effects of the drug did appear as side effects but these effects were no more severe or frequent than those of other anthelminthics. Trichlorfon is effective for treating even Schistosoma haematobium infestation (King et al., 1988). The control of botfly larvae and fleas is certainly due to the anticholinesterase and other antiesterase action of the drugs and no other mode of action is known against the susceptible nematodes. The fact that the compounds can be effective against parasites in the tissues without injuring the host is striking evidence of very great differences in suscep­ tibility to absorbed drug. The mechanism of the difference

65

is poorly understood but probably depends on the greater susceptibility of the esterases of the parasite and the greater metabolic power of the microsomal enzymes of the host. (It may be noted that systemics for plants are compounds capable of being absorbed by one part of the plant and then translocated to another part so that the plant becomes pesti­ cidal. Absorption usually occurs through the roots but may occur through the leaves or other plant organs.) Differences between Vertebrates  As a general rule, small species of warm-blooded animals eat more food than large ones in relation to their body weight. Therefore, if both kinds receive the same contaminated diet, the small species will receive a large dosage of toxicant. However, the most notable examples of species differences do not depend on differences in food intake or respiratory pat­ tern but are inherent. Sometimes the inherent difference may be explained in terms of metabolic or genetic differences. Very large differences in susceptibility are more likely among spe­ cies belonging to different phyla but may occur among spe­ cies of the same class. Norbormide is associated with a wide range of susceptibility among mammalian species. Albino Norway rats have an LD 50 of 4.3 mg/kg, whereas dogs, cats, monkeys, sheep, chickens, and ­turkeys are unaffected by sin­ gle doses at the rate of 100 mg/kg (Roszkowski et al., 1964; Roszkowski, 1965). Thus, the factor of difference for these species for norbormide is greater than 23. Even the wild strain of Norway rat is less susceptible (LD 50, 12 mg/kg). Another example of marked species susceptibility involves ducks and diazinon. Apparently the largest difference in species susceptibil­ ity that has been measured is that for the acute oral toxic­ ity of TCDD in the guinea pig and hamster; this difference may be more than 8000-fold. This large species difference was so frightening that it gave rise to the most conserva­ tive risk assessment ever conducted (0.008 pg/kg/day daily intake), because the reason for this extremely large spe­ cies difference was not understood. In the mean time it became clear that there is no such huge difference in terms of chronic toxicity and that the extraordinary difference in acute toxicity is due to differences in the way an ­herbivore (guinea pig) and a hibernator (hamster) can handle a dis­ turbance of glucose metabolism. (Rozman, 1992; Fan and Rozman, 1994) The devastating lesson from the dioxin issue for toxicology is that having understood almost any­ thing and everything about this class of compound at the cost of billions of the year 2,000 dollars has not changed the original risk assessment (based on very limited data) one iota. A cynical observer might conclude that there is an inverse relationship between toxicological investigation and “related” risk assessment. Another very large difference involves the teratogenic effects of thalidomide. Human embryos have been deformed by as little as 0.5–1.0 mg/kg taken daily by the mother for

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66

several days. At the other extreme, no injury to cat embryos was produced by a maternal dosage of 500 mg/kg/day, indi­ cating a difference of more than 1000-fold. In fact, all other species studied require a dosage greater than that which is teratogenic in some women. The rabbit responds rather uni­ formly to dosages of 30–50 mg/kg/day, whereas many strains of rats fail to respond to 4000 mg/kg/day even though a dos­ age of 50 mg/kg/day is teratogenic in a few strains (Kalter, 1965). Schumacher et al. (1968b) reported that thalidomide given orally to rats was poorly absorbed. When the drug was given intravenously at a rate of 10 mg/kg/day, malfor­ mations and resorptions were observed (Schumacher et al., 1968a). However, the basic difference remains. According to Schardein (1985) all chemicals known to be teratogenic in humans except some coumarin derivatives have been shown to be teratogenic in laboratory animals. That does not guarantee that a new compound always will be tested in a susceptible species and by an effective route. Perhaps the most diagrammatic difference in the sus­ ceptibility of species to a poison involves the use of diphac­ inone (called diphenadione as a drug) to control vampire bats that feed on cattle. When injected intraruminally at a rate of 10 mg/kg, as recommended, the compound is harm­ less to cattle but fatal to any bat that feeds within 72 hours of treatment (Mitchell, 1968). Whereas the exact degree of difference between these two species does not seem to have been measured, it clearly is substantial because the method has been in routine use for years without injury to cattle and with excellent control of vampire bats. Appleman and Feron (1986) evaluated toxicity data from 66 compounds that were tested in only two species (rats and dogs) to determine how frequently the dog provided data that were substantially different in a qualitative or quanti­ tative nature from those obtained in rats. In this evalua­ tion, the rat was highly predictive of responses in the dog. Applications of a 10-fold margin to the rat data accounted for nearly all the differences between rats and dogs. Some differences between species may be of such degree that they are essentially qualitative. Examples include the propensity of ducks to develop cataracts of the lens in response to dinitro compounds, or that of hens and humans to develop delayed but permanent paralysis in response to tri-o-cresylphosphate and some other organic phosphorus compounds. Although these are unusual situations, they emphasize how different one species can be from another. Table 1.7 shows that, for many pesticides, the factor of dif­ ference in susceptibility of the mouse, guinea pig, rabbit, and dog ranges from 0.2 to 11.8 and averages close to 1.0 in com­ parison with the susceptibility of the male rat. The factor is close to 1.0 for many anticancer agents also (Freireich et al., 1966). The fact that the species difference is usually small is confirmed by comparisons based on the kind of test that may be applied to both people and experimental animals or on information obtained in connection with accidental exposure of humans. Comparisons of the susceptibility of humans and

Table 1.7  Relative Susceptibility of Different Species to Pesticides Based on Oral LD 50 Valuesa Using the Male Rat as a Standardb Susceptibility factorc

Compound Mouse

Guinea pig

Rabbit

Dog

Chlorobenzilate

1.4

—

—

—

DDT

0.3–0.8

0.3

0.3–0.5

—

Methoxychlor

2.7–3.8

—

—

—

Lindane

1.0

0.7–0.9

0.4–1.5

—

Aldrin

0.9–1.2

1.2–1.6

0.5–1.1

0.4–0.8

Chlordane

0.8

—

1.1–3.4

—

Dieldrin

1.2

0.9–1.0

0.9–1.0

0.7–0.8

1.8–6.2

—

d

e

Endrin

—

0.5–2.7

Heptachlor

0.6–1.5

0.8–0.9

—

—

Azinphosmethyl

1.6

0.2

—

—

Chlorthion

0.7

—

—

—

Diazinon

0.9–1.4

0.3

0.8

—

Dimethoate

3.0–5.4

0.5–0.9

—

—

Dioxathion

—

—

—

1.1–11.8

Malathion

1.6–1.9

—

—

—

Methyl parathion

0.4

—

—

—

Mevinphos

0.9–1.6

—

—

—

Oxydemetonmethyl 1.0–2.5

0.3–0.6

—

—

Parathion

0.2–5.0

1.6–3.2

0.5–3.0

—

Phosphamidon

1.3

—

—

—

®

a

Data from Gaines (1960), Lehman (1965). From Hayes (1967a), by permission of the Royal Society, London. c A factor of less 1.0 indicates less susceptibility than that of the male rat; a factor greater than 1.0 indicates greater susceptibility. d Both sexes. e Approximate. b

animals are shown in Tables 1.8 and 1.9. The tables indicate, for example, that humans are more ­ susceptible than rats to lindane, about as susceptible as female rats to parathion, and distinctly less susceptible to warfarin than rats of either sex. Although in many instances in which a direct compari­ son is possible there is no marked difference, the difference appears to be about 100-fold for a few compounds and over 1000-fold for thalidomide. Furthermore, the tables by necessity present phenomena that can be studied in labo­ ratory animals. Much of the reason for conducting tests in humans is the existence of hypersensitivity, subjective responses, and other phenomena that do not lend them­ selves to study in animals.

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

67

Table 1.8  Comparison of the Susceptibility of Humans and Other Animals to Certain Pesticidesa Compound

Species (sex)

Dosage (mg/kg)

References

Largest without clinical effect

Smallest with clinical effect

Median CD 50

Smallest Largest Smallest LD 50 Uniformly with nonfatal fatal fatal serious effect

Human

—

6

10

16b

285c

—

—

—

Rat (F) Rat (M)

— 25

— —

— —

75 50

150 175

100 50

118 113

200 200

Human

—

0.4e

0.4e

0.5b

—

—

—

—

Rat (F) Rat (M) Calf

— — 2.5

— — —

— — —

— 50 5

125 125 —

75 75 —

91 87 —

— 200 —

Graeve and Herrnring (1951) Gainesd Gainesd Radeleff et al. (1955)

Chlordane

Human Human Infant Rat (F) Rat (M)

— — — 100 —

— — — 200 —

— — — — —

32b,c — 10b 300 250

— — — 550 400

— 29–57 — 350 250

— — — 430 335

— — — 600 450

Dadey and Kammer (1953) Hayes (1963) Lensky and Evans (1952) Gainesd Gainesd

Dieldrin

Human Rat (M, F)

— —

— —

— —

10 30

— 60

— 30

— 46

— —

Princi (1952) Gainesd

Endrin

Human Rat (F) Rat (M)

— — —

— — —

— — —

0.2 6 10

— 10 25

— 6 10

— 7.5 17.8

— — 30

Davies and Lewis (1956) Gainesd Gainesd

Dichlorvos

Human Rat (F) Rat (M) Cow Horse

— — — — —

— — — — —

— — — — —

51 — — 27 25

— 100 125 — —

— 37 75 — —

— 56 80 — —

— 125 150 — —

Hayes (1963) Gainesd Gainesd Tracy et al. (1960) Jackson et al. (1960)

Diazinon

Human Human Rat (M) Calf

— — — —

— — — —

— — — —

2.2f — 200 1

— 250 300 —

— — 200 10

— — 250 —

— — 350 —

Hayes (1963) Bockel (1957) Gainesd Radeleff et al. (1955)

Malathion

Human Rat (F) Rat (M) Sheep

— — 500 150

— — — 100

— — — —

— 750 1000 100

200 1250 1750 300

71 750 1000 150

— 1000 1375 —

— 1500 2000 —

Walters (1957); Paul (1960) Gainesd Gainesd Radeleff et al. (1955)

Parathion

Human

—

—

—

—

6.4

2.0

—

13

Child Rat (F) Rat (M) Calf Sheep Steer

— — 5.0 — 50 25

— — — — — —

— — — — — —

— 1 10 0.5 — —

— 4.5 20 — 75 —

0.1 3.0 10 1.5 20 —

— 3.6 13.0 — — —

— 5 30 — — —

Goldblatt (1950); Hayes (1963) Kanagaratnam et al. (1960) Gainesd Gainesd Radeleff et al. (1955) Radeleff et al. (1955) Radeleff et al. (1955)

Human

—

0.05

—

3.5

—

—

—

—

Rat (M)

—

—

—

—

—

1.0

1.05

—

DDT

Lindane

TEPP

a

From Hayes (1967a), by permission of the Royal Society, London. All doses are single and oral unless otherwise noted. Convulsions. c Part of does vomited. d Based partly on published papers (Gaines, 1960, 1969) and partly on the original data from which the papers were drawn. e Three times a day for 3 days; highly dispersed formulations. f Dermal. b

Garrett (1947); Hsieh (1954); Neal et al. (1946); Velbinger (1947); Hayes (1959) Gainesd Gainesd

Grob and Harvey (1949) Grob et al. (1950) Gainesd Gainesd

Hayes’ Handbook of Pesticide Toxicology

68

Table 1.9  Comparison of the Susceptibility of Humans and Other Animals to Repeated Doses of Certain Pesticidesa Compound

Species (sex)

Dosage (mg/ kg/day)

Duration (days)

Resultsb

Reference

DDT

Human Rat (M, F)

0.5 0.24

600 161

Increased storage; no clinical effect Histopathological changes of the liver

Hayes et al. (1956) Laug et al. (1950)

Methoxychlor

Human Rat (M, F)

2 4.87

56 750

No effect No effect level

Stein et al. (1965) Lehman (1965)

Demeton

Human Rat (F) Rat (F) Rat (F)

0.05 0.05 0.14 0.24

24 112 112 66

Dog (M, F) Dog (M, F)

0.025 0.047

168 168

15% reduction of plasma ChE only No significant depressions of ChE 30% inhibition of ChE 60% reduction of plasma; 40% reduction of RBC ChE No significant depression of ChE Significant depression of ChE

Moeller and Rider (1962b) Barnes and Denz (1954) Barnes and Denz (1954) T. B. Gaines, unpublished results (1962) Frawley and Fuyat (1957) Frawley and Fuyat (1957)

Human Human Rat (F)

0.002 0.0034 0.024

70 70 28

Edson (1964) Edson (1964) Edson (1964)

Rat (F)

0.095

28

Rat (F)

0.475

28

No effect on ChE 25% reduction of whole blood ChE About 50% inhibition of RBC ChE; no effect on plasma ChE 75% reduction of RBC ChE; 25% reduction of plasma ChE Almost complete inhibition of RBC ChE; 75% reduction of plasma ChE

Human Human

0.05–0.075 0.15

59 28

Frawley et al. (1963) Frawley et al. (1963)

Rat (M, F) Rat (M, F)

0.22 0.78

91 91

Dog (M, F)

0.25

12

Dog (M, F)

0.8

12

No inhibition of RBC and plasma ChE Slight inhibition of plasma ChE; no effect on RBC ChE No significant effect on ChE Significant reduction of RBC and plasma ChE Marked effect on plasma ChE; no effect on RBC ChE Marked effect on plasma ChE; no effect on RBC ChE

Human

0.34

56

Moeller and Rider (1962a)

Rat (F)

3.2

90

Rat (M)

4.5

730

Maximal reduction of 25% plasma and RBC ChE 29% reduction in RBC and no reduction of plasma ChE on 30th day; recovery by 90th day 10–30% inhibition of plasma and RBC ChE

0.1 0.94

24 84

15% reduction of plasma ChE only Significant depression of plasma and RBC ChE

Moeller and Rider (1962b) Williams et al. (1959)

Human

0.1

42

Edson (1964)

Rat (F)

0.07

90

33% reduction of whole blood ChE; 16% inhibition of RBC ChE; 37% inhibition of plasma ChE No effect

Rat (F)

0.26

84

Edson (1964)

Rat (F)

0.35

90

Dog Pig

0.047 4.0

168 49

80% reduction of RBC ChE; slight inhibition of plasma ChE 37% reduction of plasma and 44% reduction of RBC ChE 60% inhibition of plasma ChE 80% inhibition of RBC ChE; no inhibition of plasma ChE

Human Human

0.014 0.06

44 60

Edson (1964) Edson (1964)

Rat (M, F)

0.045

112

25% reduction of blood ChE 77% inhibition of RBCD ChE; 50% inhibition of plasma ChE Substantial reduction of ChE; no effect on plasma ChE

Dimefox

Dioxathion

Malathion

Methyl parathion Human Dog (M, F) Paration

Schradan

Edson (1964) Edson (1964)

Frawley et al. (1963) Frawley et al. (1963) Frawley et al. (1963) Frawley et al. (1963)

T. B. Gaines, unpublished results (1968) Hazleton and Holland (1953)

T. B. Gaines, unpublished results (1968)

T. B. Gaines, unpublished results (1968) Frawley and Fuyat (1957) Edson (1964)

Edson (1964)

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

69

Table 1.9  (Continued) Compound

Arsenic trioxide

Warfarin

2,4-D

Species (sex)

Dosage (mg/ kg/day)

Duration (days)

Resultsb

Reference

Rat (M)

0.22

14–85

Complete inhibition of RBC ChE

Edson (1964)

pig (F)

0.1

102

55% inhibition of RBC ChE; slight reduction of plasma ChE

Edson (1964)

Human

0.44

?

Frequent mild poisoning

Sollmann (1957)

Sheep

10

—

Tolerated without symptons

Reeves (1925)

Horse

4.7

—

Tolerated without symptons

Reeves (1925)

Human Human

0.14 0.29–1.45

Indefinite 15

Friedman (1959) Lange and Terveer (1954)

Human

1.7

6

Human

0.83–2.06

15

Maintenance therapeutic dose Hemorrhage in 12 people (4–70 yr) followed by recovery Hemorrhage in 22 yr man followed by recovery Fatal to boy (19 yr) and girl (3 yr)

Rat (M, F)

0.08

40

Killed 5 of 10

Hayes and Gaines (1959)

Rat (M, F)

0.39

15

Killed 10 of 10 rats

Hayes and Gaines (1959)

Human Human

14–37 66

18 1

Tolerated Coma, hyporeflexia, incontinence

Seabury (1963) Seabury (1963)

Rat (F)

15

112

Tolerated

Hill and Carlisle (1947)

Dog

9

84

Tolerated

Drill and Hiratzka (1953)

Holmes and Love (1952) Lange and Terveer (1954)

a

From Hayes (1967a), by permission of the Royal Society, London. All doses are oral unless otherwise noted. RBC, red blood cells; ChE, cholinesterase.

b

Species differences may involve any parameter. McCully et al. (1968) found as much as 40-fold difference in tissue levels of DDT following a single oral dose at the rate of 10 mg/kg to rats, sheep, chickens, rabbits, and guinea pigs. The greatest difference for three routes (oral, intraperitoneal, and intramuscular) in any one species was five-fold. Species differences in the effect of inhaled chemi­ cals on pulmonary function complicate the use of con­ centration in air as an indicator of dosage. Studies with methyl bromide (Medinsky et al., 1985) and formaldehyde (Chang et al., 1981) revealed significant species differences in the response of pulmonary function to these two inhaled chemicals. Changes in pulmonary function were such that dosage was not simply a function of concentration, time of exposure, and normal minute volume of the species. Thus, without knowledge of the effect on pulmonary function, extrapolation between species is particularly difficult. Certain animals lend themselves to testing for a limited number of specific forms of toxicity. Thus, hens are used to screen for possible neurotoxicity of organic phosphorus compounds, not only because they are highly susceptible to this injury, but because this susceptibility seems to resemble

that of humans. In a similar way, ducks are used to test the tendency of dinitro compounds to cause cataracts. The ideal scheme would include a species of experimen­ tal animal resembling humans so closely in ­ susceptibility to poisons that any differences would be unimportant. Unfortunately, no such animal has been identified or is likely to exist. Monkeys and apes have been suggested. They are valuable for special purposes, but there is no con­ vincing evidence that their average value is greater than that of any other laboratory animal. In some instances they are distinctly inferior to other animals. For example, when fed DDT, rhesus monkeys metabolize little or none to DDE, although both rats and humans form this compound readily (Durham et al., 1963). Variation between species must be considered every time a new compound that has been properly tested in ani­ mals is used for the first time. It is frequently suggested that the tests be made in a large number of species. If the results in the different animals are similar, it is likely that human response to the compound will not be greatly differ­ ent. If, however, there is a wide variation in the response of different species, then conservatism forces us to suppose,

70

until there is direct evidence to the contrary, that humans may be at least as sensitive as the most susceptible species. No matter what the pattern of response in experimental animals is, the ultimate tests must be in humans. It is best that such tests be carried out under circumstances permitting scientific observations. Strain Differences  Strain differences in background disease and susceptibility to toxicants has influenced the selection of strains of laboratory animals for use in routine tests. For example, regarding models for carcinogenicity testing, strains of animals, even though in common use, have differences, which are problematic. Life spans differ significantly from one strain to another and must be taken into account in select­ ing the duration of chronic studies. Sprague-Dawley rats are characterized by a fairly high incidence of mammary tumors, and the life span of the male rat is commonly limited by the background incidence of renal disease. The F344 rat, though widely used in carcinogenesis studies, has an extremely high occurrence of interstitial cell tumors of the testis, preclud­ ing the use of the testis as an organ of evaluation of testicu­ lar tumors of that cell type. The B6C3F1 mouse, particularly the male, has a high and variable incidence of liver tumors (Haseman et al., 1984). There is no single strain or species of laboratory animal that is clearly most predictive of chronic toxicity or carcinogenicity for humans. That statement can also be made for most other end points of toxicity. Thus, acceptance of one or multiple species of animal for testing must recognize the limitations of each species and strain, and interpretation of results must be made accordingly. Individual Differences  Individual differences are appar­ ent in every toxicological test, including those carried out in people. A paper by Gaines (1969), in which he reports the acute oral toxicity of pesticides, shows that for 69 com­ pounds the LD 50 value for male rats was 1.20–7.14-fold greater than the corresponding LD 01 value. The average factor of difference was 2.42. The corresponding factors of difference for the dermal toxicity of 42 pesticides were 1.37–14.93 with an average of 3.00. In other words, judged in this way, individual variation, although very real, is usu­ ally relatively small. In studies of storage and excretion, the greatest individual average excretion of malathion-derived material differed from the smallest individual average excretion at the same dosage by factors of only 2.2–8.7 for different groups of people (Hayes et al., 1960). Thus, the degree of difference was relatively constant in tests carried out at different dosage levels or at different times. A ­similar observation was made regarding the storage of DDT and the excretion of DDA in humans. In separate tests, the max­ imal storage of DDT was 1.3–5.9 times the minimal stor­ age at the same dosage level. For a single dosage group, the maximal rate of excretion of DDA by one man in any one day was 18.0 times greater than his own minimal rate, and the difference between the lowest minimum and highest

Hayes’ Handbook of Pesticide Toxicology

maximum within the group was a factor of 21.5 (Hayes et al., 1971). In all of these tests, the relative constancy of one individual compared with another was noted. Although individual differences may be described in statistical terms, physiological understanding of these dif­ ferences is lacking almost entirely. If a population is suf­ ficiently heterozygous, the differences between individuals may depend on their genetic diversity. However, individual differences persist to some degree in a homozygous popu­ lation (see Section 1.3.1.4 for discussion on Gaussian dis­ tribution). This is illustrated by the failure of the LD 50 values of four pesticides in a particular population of mice to change in the course of l2 or more generations even though each succeeding generation was bred from mice that had survived an LD 50 dose (Guthrie et al., 1971). Sex  Chiefly, because of its convenience, the rat is used more than any other species for studies in toxicology. The rat also has the apparent distinction of showing more varia­ tion between the sexes in its response to chemicals than any other species. This fact may have led to more concern than is justified regarding possible differences in the sus­ ceptibility of men and women to chemicals. In any event, calculations from data provided by Gaines (1960, 1969) for the oral toxicity of 69 pesticides showed that the dif­ ference in the oral LD 50 for male and female rats ranged from 0.21 (indicating greater susceptibility of the female) to 4.62 (indicating greater susceptibility of the male), and averaged 0.94. The corresponding factors for the dermal toxicity of 37 pesticides were from 0.11 to 2.93, with an average of 0.81 (Hayes, 1967a). The differences in the sus­ ceptibility of male and female rats are associated to a large degree with differences in their liver microsomal enzymes. In contrast to the situation in rats and to a lesser degree in other rodents, significant differences between the sexes of other species in their susceptibility to poisons usually have not been reported. Such differences were looked for but not found in studies of the storage of DDT in monkeys (Durham et al., 1963; Ortega et al., 1956). Such differences between men and women are small or lacking entirely. Pregnancy  Susceptibility to a particular compound may be either greater, less, or identical in pregnant females than in nonpregnant ones of the same strain and age. For exam­ ple, pregnancy exaggerates the danger of anticoagulants but reduces the danger of paraquat. For some differences such as susceptibility to anticoagulants the reason for the difference is clear. In most instances the reason is obscure. In a systematic study of 19 drugs, given by different routes, pregnant mice were more susceptible than nonpregnant ones by factors ranging from 0.74 to 14.55 and averaging 1.90, or 1.27 if the single high value is excluded (Beliles, 1972). Lactating rats consume approximately 3-fold (Hayes, 1976) or 2.5–3-fold (Yang et al., l984a) more feed than the same rats before or after lactation; thus lactating rats

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

are subject to a marked increase in the dosage of all com­ pounds in their diet. The extent to which this is true of other species apparently is not documented. Other Endocrines  There is considerable evidence that the pituitary adrenal axis may be influenced by photoperiodic­ ity and in turn may influence susceptibility to toxicants. Lipsett (1983) has reviewed the relationship between hormones and cancer. Sex and other hormones usually have either pulsatile release or some cyclicity (diurnal or circadian cycles) in their secretion. In either event, an additional timescale needs to be incorporated into studying these phenomena in the context of toxicity in addition to the three independent timescales discussed earlier. Gender differences will become manifest in toxicology only when the aforementioned timescale becomes rate-determining (or-limiting). Pregnancy introduces still another timescale with known consequences for altered hormonal timescales for a limited period in an individual’s life. Age  Children and young animals are often more suscep­ tible than adults to poisons in food. The most common rea­ son is that children and other young animals eat more than adults in proportion to their weight. Thus, when given the same contaminated food, young animals receive a higher dosage of toxicant. The relationship for rats is shown in Fig. 1.19. Although the figure is based on DDT, it applies equally to any compound that does not cause a reduction of food intake.

71

However, other factors may be involved. It is now well known that a number of drugs are poorly metabolized by infants, particularly those born prematurely (Fouts and Adamson, 1959). Although it is seldom possible to quan­ titate the difference, it is clear that a dosage of some drugs easily tolerated by human adults may lead to severe illness or even death in very young children. Calves and sometimes lambs are markedly more sus­ ceptible than adult cattle or sheep to sprays or dips of chlordane, dieldrin, and lindane (Radeleff, 1970). Systemic study of drugs (Hoppe et al., 1965; Yeary and Benish, 1965) and pesticides (Brodeur and DuBois, 1963; Gaines and Linder, 1986; Lu et al., 1965) indicates that new­ born animals are generally more susceptible than adults of the same species regardless of route of administration. However, the differences tend to be small and there are some exceptions in which the newborns are actually less susceptible than adults (Brodeur and DuBois, 1963; Gaines and Linder, 1986; Hoppe et al., 1965). As reviewed by Durham (1969), the difference, no matter what its direction, can often be explained in terms of the recognized activity of microsomal enzymes in activating or deactivating the chemical in question. Other factors that may be of importance are renal function and membrane permeability. Fortunately, compounds of similar pharmaco­ logical action tend to show similar differences in their tox­ icity to young adult animals (Yeary and Benish, 1965). In studying drugs, Yeary and Benish found that newborn rats were 0.6–10.0 times more susceptible than adults. Hoppe et al. (1965) found a similar range of 0.7–6.2. Goldenthal

DDT intake (mg/kg/day)

70 60 50 800 ppm

40 30

800 ppm 600 ppm 600 ppm 400 ppm

20

400 ppm

10

200 ppm 100 ppm

0

1

2

4

6 Months

8

10

12

FIGURE 1.19 Calculated DDT intake (mg/kg body weight) in rats receiving various levels of DDT in the die (male, —; female,---). From Fitzhugh and Nelson (1947), by permission of the Williams & Wilkins Co., Baltimore.

72

(1971) studied a much larger number of drugs and a few other compounds (290 in all) and reported a much wider range of factors of differences: 0.2–750. However, by omitting only one low factor and eight high ones, the range was narrowed to 0.1–20. The geometric mean of all of the factors (no exceptions) was 2.78. In their studies of 15 organic phosphorus insecticides or defoliants, Brodeur and DuBois (1963) found that wean­ ling male rats were 0.2–4.1 times more susceptible than adults, with a mean variation of 1.8 DDT is also less toxic to infant rats than to adults. Age-related differences in susceptibility to carcino­ gens have not been given extensive attention, but there is increasing concern about the potential for transplacental carcinogenesis (carcinogenic effects associated with in utero exposure). Few chemicals are known to be carci­ nogenic by in utero exposure only (and not from lifetime or long-term exposure after weaning); however, in cer­ tain cases the profile of tumors is known to be different between the two exposure patterns, for instance, in the case of the carcinogenic effects of diethylstilbestrol in humans. This area has been summarized and reviewed by Rice (1984). Warzok et al. (1980) found differences in carcino­ genicity after transplacental and postnatal administration of drugs, pesticides, or their metabolites. Comparative studies of the carcinogenic activity of procarbazine, methylphenyl nitrosourea, and ethylenethiourea after transplacental and postnatal administration showed marked differences in the frequency and spectrum of tumors induced. For example, administration of procarbazine transplacentally and post­ natally resulted in a much higher production of rats with tumors than administration of procarbazine only transpla­ centally or only postnatally. Age-related differences in physiology and metabolism probably account for many of the age-related differences in manifestations of toxicity to chemicals. Borghoff and Birnbaum (1985) identified clear age-related changes in glucuronidation and deglucuronidation that depended on the chemical substrate as well as the tissue as a function of age. Also, significant physiologic changes, such as dis­ tribution of fat, may account for certain age-related differ­ ences in metabolism and toxicity as shown by Yang et al. (1984b) in rats dosed for 2 years with ethylenediamine. The definition of toxicity (Section 1.2) implies that aging is a toxicological phenomenon amounting to accu­ mulation of injury in an organism over short or long periods of time. This accumulation of injury is due to ther­ modynamics, which means that there are no truly revers­ ible phenomena in nature, even though some processes might get close to it. Therefore, it must be understood that those people who maintain that modern medicine will be capable of prolonging life almost indefinitely are latter day protagonists of some sort of perpetuum mobile, which in terms of science is nonsense. Even replacing organs buys only a short reprieve in old age, because second, third, and

Hayes’ Handbook of Pesticide Toxicology

subsequent breakdown points are ever closer to each other making the time gained shorter and shorter at exponen­ tially increasing costs. A toxic insult may be no different in a young than in an old individual if the damage can be recovered from very rapidly (kinetic or dynamic recovery). However, the more irreversible (due to either kinetics or dynamics) an injury is, the larger will be the c  t contribution to the eventual demise of an organism. Therefore, concern about and pro­ tection from exposure of fetuses, neonates, and the young to chemicals with long kinetic or dynamic half-lives is a highly legitimate goal and one of the most important tasks of modern toxicology. However, claiming the need for a 10-fold safety factor for all chemicals to protect this popu­ lation from significant toxic insult can only originate out from ignorance of the science of toxicology. Nutrition­  The relationship between spontaneous or chem­ ically induced carcinogenesis and nutrition has been stud­ ied extensively and has been the subject of several reviews (Campbell, 1979; Clayson, 1975; Everett, 1984; Rao et al., 1987; and a symposium introduced by Omaye, 1986). How­ ever, the subject is beyond the scope of this volume except as it may involve pesticides. General Nutritional Condition  Apparently only extremes of general nutrition have produced observable alterations in the toxicity of pesticides. As reviewed elsewhere (Hayes, 1959), various mammals and even fish are relatively resis­ tant to poisoning by DDT if they are fat rather than thin. The same result has been produced with dieldrin and lin­ dane under experimental conditions (Barnes and Heath, 1964; Geyer et al., 1993). Other factors may be involved, but certainly distribution of the insecticide to adipose tis­ sue tends to reduce the concentration of the insecticide at the site of action and thus protecting the organism. Paired feedings may be used to distinguish those effects of a toxicant secondary to reduced intake of food (Weber et al., 1991). Effect of Starvation  If DDT is stored in body fat in suf­ ficient concentration, rapid mobilization of the fat through starvation may lead to poisoning (Fitzhugh and Nelson, 1947). There is an increase in the concentration of poi­ son in the small amount of fat remaining and, by the same token, in all tissues of the body (Dale et al., 1962). During mobilization of DDT, excretion is increased by a factor of about 1.4 but the increase is insufficient to prevent poi­ soning in some rats. Dale and associates pointed out that starvation is unlikely to precipitate poisoning by DDT in humans because even people with heavy occupational exposure to the compound do not store enough of it to pro­ duce the effect and because the metabolism of humans is inherently slower than that of rats so that humans cannot starve as quickly.

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

Some other chlorinated hydrocarbon insecticides are excreted more efficiently than DDT. For example, Heath and Vandekar (1964) found that the average excretion of dieldrin in rats was relatively rapid (5%/day), and that it was more than doubled following a few days of starvation. Therefore, it is not astonishing that it was not possible to precipitate dieldrin poisoning in rats by starving them after they had been fed for 7–18 months at dietary levels up to 15 ppm (Treon and Cleveland, 1955). The toxicological effect of weight loss associated with infection may be indistinguishable from those associated with starvation (see this Section further below). Quantity of Dietary Protein  In addition to the action involving storage just discussed, nutrition may influence toxicity through metabolism promoted by the liver micro­ somal enzymes. Murphy and DuBois (1957) showed that male rats maintained on a protein-free diet for 4 weeks had only 24% of the microsomal enzyme activity of normal rats. Also, the liver enzymes of rats that did not receive protein could not be induced by compounds that ordinarily would stimulate these enzymes. There is great variation in the influence of protein mal­ nutrition on susceptibility to acute poisoning by different compounds. This variation may depend on (a) the net effect of biotransformation, whether detoxification or intoxica­ tion (metabolic activation); (b) differences in the relative contribution of biotransformation to toxicity; (c) the ability of some compounds to cause anorexia or other disruption of nutrition; or (d) other mechanisms. As an example of a difference in the net effect of bio­ transformation, it may be recalled that the toxicity of aflatoxin, which is detoxified by the liver, is increased by protein deficiency whereas the toxicity of carbon tetrachlo­ ride, which is rendered toxic by metabolism, is decreased greatly by protein deficiency (McLean and McLean, 1969). The LD 50 of DDT and the associated clinicopathologi­ cal effects showed only slight variation among rats previ­ ously maintained for 4 weeks on ordinary laboratory feed and rats fed a synthetic diet containing either normal protein (27% casein) or deficient protein (8% casein) (Boyd and De Castro, 1968). Even after a diet containing no protein, the toxicity of DDT was increased only fourfold (Boyd and Krijnen, 1969b). By extreme contrast, the acute toxicity of captan was increased 2100-fold in rats maintained without protein compared with those fed normal protein (Krijnen and Boyd, 1970). The results for DDT, captan, and a num­ ber of other pesticides are summarized in Table 1.10. It may be seen that even a very great increase in protein is without important effect on susceptibility to pesticides. The same is true of a reduction of protein to only one-third of the nor­ mal level. However, when protein restriction is severe and especially when it is complete, susceptibility to some pesti­ cides is increased dramatically. It must be recalled that rats that have been maintained without protein for 28 days after

73

Table 1.10  Estimates of the Increase in the Acute Toxicity of Certain Pesticides in Albino Rats as Related to the Concentration of Protein in Their Diet during 28 Days from Weaning until Dosinga Agent

Percentage in diet

Reference

0.0

3.5

9.0

26.0

81.0

Captan

2,100

26.3

1.2

1.0

2.4

Krijnen and Boyd (1970)

Carbaryl

8.6

6.5

1.1

1.0

1.0

Boyd and Krijnen (1969a)

CIPC

8.7

4.0

1.7

1.0

—

Diazinon

7.4

1.9

1.8

1.0

2.0

Boyd et al. (1969a)

DDT

4.0

2.9

1.5

1.0

3.7

Boyd and Krijnen (1969b)

Endosulfan 20.0

4.3

1.8

1.0

1.0

Boyd et al. (1970)

Lindane

12.3

1.9

1.0

1.0

1.8

Boyd et al. (1969b)

Monuron

11.5

3.0

1.8

1.0

—

Boyd and Dobos (1969)

1.0

—

Boyd and Taylor (1971)

Toxaphene —

3.7

Boyd and Carsky (1969)

a From Boyd et al. (1970), by permission of the American Medical Society.

weaning, weigh about 30% less than they did when placed on the diet. Two-thirds of these rats die in the first 3 days after their food is withdrawn even though no chemical is administered. It is little wonder that their susceptibility to compounds that cause anorexia is striking. Other compounds that have been studied in relation to protein deprivation include carbanolate, parathion, chlor­ dane (Casterline and Williams, 1969), dieldrin (Lee et al., 1964), and TCDD (Muzi et al., 1987). In a thorough review of diet and toxicity, McLean and McLean (1969) emphasized the opposite effects of protein deficiency on the toxicity of compounds that are detoxi­ fied and those that are made toxic by biotransformation, especially in those instances in which the site of biotrans­ formation is also the site of toxic injury. The reviewers also pointed out that reversal of one aspect of deficiency (such as the induction of microsomal enzymes by a foreign com­ pound or by a component of natural diets in animals with borderline protein deficiency) may reverse the entire effect of diet on toxicity. Although there is evidence that malnour­ ished people are unduly susceptible to infection, there is no

74

clear evidence that the cell’s general ability to withstand change and trauma is altered by malnutrition. The relation of nutrition to toxicity must be determined separately for each compound and, under practical conditions, other fac­ tors must be taken into account. Quality of Dietary Protein  The acute oral toxicity of heptachlor was found to be 1.6–2.1 times greater in rats that were pair-fed casein than in those that were fed glu­ ten, regardless of whether protein constituted 10 or 18% of the diet. The difference was less or even reversed when the casein diet was fed ad libitum and weight gain was greater (Webb and Miranda, 1973). Gluten is an incomplete protein that reduces food intake and permits only an abnormally small increase in body weight of rats that consume it ad libitum as their only source of protein. It seems likely that the lower toxicity of heptacholor in rats fed gluten depends on limited conversion of the compound to heptachlor epoxide as a result of limited activity of the microsomal enzymes of the liver. On the other hand, the even greater protection offered by normal intake of high quality protein may result from the presence of normal fat deposits and the sequestering of both heptachlor and its epoxide in fat. Effects of Fat  Dietary fat has been studied less than dietary protein in relation to pesticides. However, Purshottam and Srivastave (1984) found that a high-fat diet significantly protected against mortality from an indirect inhibitor of cholinesterase (parathion) but not from a direct inhibitor (dichlorvos). In contrast, a high fat diet increased the acute toxicity of TCDD compared to high-carbohydrate-fed rats (Muzzi et al., 1987). Most studies of fat that are potentially of toxicological interest involve carcinogenesis. However, because the findings apply to spontaneous as well as to induced tumors, the rele­ vance to toxicology may be obscure. Briefly, breast and uter­ ine cancer are more frequent in obese women. The possible hormonal basis of this relationship has been discussed (Lipsett, 1983). In fact, several tumors are more common in people who are overweight (Doll and Peto, 1981). An increase in dietary fat-or, in fact, any variable leading to an increase of many spontaneous tumors in experimental animals (Rao et al., 1987; Ross et al., 1983) – increases the yield of tumors induced by chemicals in experimental animals (Bin et al., 1983; Chan and Cohen, 1974; Kollmorgen et al., 1981; O’Connor et al., 1985; Tannenbaum, 1940). There is some question whether dietary fat has a specific effect or merely contributes calories inasmuch as restricted intake of a particular diet increases lon­ gevity and decreases the incidence of tumors (Boissonneault et al., 1986; Conybeare, 1980; Rehm et al., 1985). In fact, the restriction of diet may be protective even if it is imposed only for several weeks after weaning (Ross and Bras, 1971). On the other hand, for at least some tumors an increase in inci­ dence depends on the specific composition of a fat regardless of its concentration in the diet. Thus, dietary levels of 0.3, 1,

Hayes’ Handbook of Pesticide Toxicology

or 10% corn oil (which contains linoleate) resulted in more mammary tumors than a dietary level of 10% corn oil from which the linoleate had been eliminated by hydrogenation (Abraham et al., 1984). Also, a control group of rats given biweekly 4 ml/kg of corn oil lived significantly longer than ad libitum-fed controls with reduced tumor incidence (Rozman et al., 2005). In designing experiments, it is important to note that rats given oil by gavage can have a threefold greater caloric intake than untreated controls (Kraft, 1983). Miscellaneous Nutritional Effects  Deficiency of any essential trace element is injurious in itself. However, a borderline deficiency may predispose to injury by a toxicant. Furthermore, there may be an interaction in the metabolism of trace elements whether essential or not. For example, Brinkman and Miller (1961) found that rats fed molybdenum gained less weight and had lower hemoglo­ bin levels if they were kept in galvanized cages instead of stainless steel cages. Similar effects were produced by increasing the zinc content of the diet of rats fed molybde­ num and kept in stainless steel cages. Isolation and Crowding  Either isolation or crowding may influence the behavior, biochemistry, and morphology of animals. Rodents have been most studied in this regard but it seems unlikely that nonrodents are immune. Although very few drugs have been studied in this way, enough work has been done to show that either isolation or crowding has a dramatic effect on the susceptibility of some strains to certain drugs, but little or no effect on their susceptibility to others. Animals are caged separately in most tests of toxic­ ity. This practice facilitates observation of each animal and permits measurement of individual food intake and collection of individual samples of excreta for analysis. It has been suggested (Hatch et al., 1963) that the results of tests on isolated rats do not reflect the functioning of nor­ mal animals. It is true that many wild rodents tend to live in small groups and that common laboratory rodents will cluster if permitted to do so. Consideration of isolation and crowding might be crucial in the study of a rodenticide from the standpoint of rodent control. The ultimate objec­ tive of most toxicity testing is the safety of humans, not that of rodents. However, in all tests of toxicity, there is a clear need to keep in mind the possible effects of isolation and crowding. Differences in the handling of animals may lead to marked differences in the results of tests in different laboratories or in the same laboratory at different times. Physiological Effects of Isolation  The effects of isola­ tion may depend on sex, strain, and duration of isolation (Wiberg and Grice, 1965). When these factors are held con­ stant, the effects may vary depending on the past grouping history of each animal (Thiessen, 1963). In other words,

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

isolation changes the susceptibility of animals to crowd­ ing. Isolated mice had relatively heavier testes and showed much less locomotion in a standard test than mice held in groups of 5 each. However, when previously isolated mice were placed in groups of 10 each, they showed increased fighting, diminution of testes weight, and a higher level of locomotor activity than mice that had been in groups of 5 before being placed in groups of 10. According to a review by Hatch et al. (1963), isolation of rats or mice for 10 days or less may produce lowered resistance to stress, lower food consumption and weight gain, and smaller adrenals, in comparison with animals held in groups of two or more. Isolation longer than a month may produce the opposite effects, including greater food consumption and a tendency toward larger adrenals. In addition, lower weights of the thyroid, thymus, spleen, and ovary, an increase in oxygen consumption, and abso­ lute leukopenia and eosinopenia have been observed. Other changes have been reported less commonly. The authors interpreted their own findings and those of others as indi­ cating that isolation produces an endocrinopathy probably involving the adrenal cortex. Effects of Isolation on Susceptibility to Chemicals Isolation may have a marked effect on the reaction of rodents to some chemicals. In other instances, isolation may change the threshold of susceptibility but have little effect on the LD 50 level (Wiberg and Grice, 1965). A very dramatic effect of isolation and its duration on susceptibil­ ity to a drug is shown in Fig. 1.20. Not shown by the figure is the fact that rats conditioned by isolation for 3 months remained highly susceptible to isoprenaline even after they had been regrouped for a week (Balazs et al., 1962). Perhaps it is the conditioning produced by isolation rather than isolation itself which is of greatest or most

75

frequent importance in influencing the action of chemi­ cals. It was shown very early (Gunn and Gurd, 1940) that -amphetamine is more toxic for aggregated than for iso­ lated mice. This result has been confirmed many times. Thus, although the susceptibility of pregrouped mice was increased by grouping as opposed to isolation immediately after dosing (mice died in an average of 53 minutes when grouped, but survived 69 minutes when isolated), the suscep­ tibility of preisolated mice was increased drastically by group­ ing but was also inherently greater (mice died in an average of 14 minutes when grouped after preisolation but survived 51 minutes when isolated) (Welch and Welch, 1966). At least in connection with stimulants of the central nervous system, the effects of aggregation may be due to hyperpyrexia associated with increased motor activ­ ity resulting from greater response to external stimuli (Peterson and Hardinge, 1967). It is impossible to discuss here the varied and complex differences in brain and adrenal catecholamines that have been shown to depend on group density. Enough has been said, however, to emphasize the importance of carrying out toxicological experiments under standardized conditions or of varying the conditions knowingly. Effect of Crowding  Crowding is not merely the absence of isolation but a deviation from the norm in the opposite direction. It can cause striking clinical injury and social disintegration, at least in some species. However, there are distinct differences even between rats and mice (Chévedoff et al., 1980; Klir et al., 1984). The phenomenon has been studied mostly in relation to population control. In a review of this aspect, Christian and Davis (1964) concluded that excess population density leads to increased aggressiveness and other forms of competition and thus (through an endo­ crine feedback mechanism involving pituitary-­adrenocortical

900 800

LD 50 (mg/kg)

700 600 500 400 300 200 100 0

0

1

2

3

4

5 6 7 8 Isolation (weeks)

9

10

11

12

Figure 1.20  Effect of isolation on susceptibility of rats to isoproterenol. Based on data from Hatch et al. (1963).

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activity and inhibition of reproduction) to regulation and limitation of population growth. This mechanism has been demonstrated for some rodents, lagomorphs, and deer, and it may apply to other mammals. According to this concept, other factors such as disease, predation, and weather limit populations occasionally, but the feedback mechanism remains as a safety device to prevent destruction of the envi­ ronment and consequent extinction. A tranquilizer can reduce aggression in a population and raise the limit at which the growth curve reaches equi­ librium. After three populations of house mice had become crowded and aggression and reproduction had leveled off, chlorpromazine was added to the diet of two of the popula­ tions at a concentration of 750 ppm. Although this concen­ tration of chlorpromazine slightly reduced the reproduction of individual pairs of mice tested separately, it decreased aggression and increased breeding success of the crowded mice. Population growth was renewed while the drug was being administered. When chlorpromazine was removed from one of the treated populations, the rate of aggres­ sion increased and the number of mice declined. The third population, which served as a control, declined slightly but probably not significantly while the other two increased under the influence of chlorpromazine (Vessey, 1967). Although crowding (in the sense used in population dynamics) is not likely to occur in a toxicology labora­ tory, population density undoubtedly has a bearing on the practical use of rodenticides and other poisons to control pests. Other Social or Psychological Factors  Anything that disturbs an animal may influence its physiological reac­ tions and thus possibly change its reactions to foreign chemicals. In some instances, disturbance may have sev­ eral components and it may be difficult to determine their relative importance. For example, each visit of the inves­ tigator or attendant to the animal room involves auditory, visual, olfactory, and sometimes tactile stimuli that ani­ mals can detect. Their responses may be unconditioned or conditioned by previous experience. The mere placing of an animal in a cage that differs in shape, area, material, or bedding may influence behavior. Changing the shape of a cage may cause mice to produce wet feces and would interfere with the testing of diuretics or purgatives (d’Arcy, 1962). Chance (1947) reported that by using larger cages he could reduce to about half the toxicity of amphetamine and ephedrine to individually caged mice. Other effects of caging were reviewed by Chance and Mackintosh (1962). Factors that may be influenced by the location of a cage in the animal room include lighting, temperature, and ven­ tilation. Any influence of the cage itself or of its location is spoken of as a “cage effect.” To cancel out such potential effects, investigators may employ random assignment of cages and periodic rotation of their locations. However, at least in regard to tumor incidence, Haseman (1988) reported

Hayes’ Handbook of Pesticide Toxicology

that among 79 dosed groups of mice the occurrence of appar­ ent cage effects agreed closely with that expected by chance. Audiogenic seizures are a dramatic, specialized response of some species to certain frequencies and intensities of noise. Approximately 36% of normal Sherman strain rats had seizures in response to intense noise from an electric bell, but the response rose to about 80% in rats receiving dieldrin at a dietary rate of 25 ppm. Although a number of compounds have been studied in this regard, the relation­ ship between response and the tone, intensity, and pulse fre­ quency of sound apparently has received little attention. Disease  Few studies have been made of the relationship between the toxicity of chemicals and the occurrence of disease of other causes. A few exceptions are clearly rec­ ognized; for example, silicosis predisposes to tuberculosis. A dosage-response relationship appears to hold, for there is no evidence that inhalation of silica insufficient to cause silicosis has any effect on the occurrence of tuberculosis. In one instance, a laboratory using pathogen-free rats con­ sistently found higher LD 50 values for a series of test com­ pounds than did other laboratories using normal rats of the same strain in a prearranged study (Weil and Wright, 1967). Hayes (1982) reported that tube feeding of rats with lar­ vae of Trichinella spiralis at the rate of 20 larvae per gram of body weight produced a marked temporary decrease in food intake and a corresponding loss of body weight of about 60 g. The loss occurred over a period of 10 days in rats receiving no other treatment, but continued at a much slower rate for another 6 days in rats that previously had received DDT at a dietary level of 200 ppm (8.5 mg/kg/day for males and 10.5 mg/kg/day for females) for 359 days before infes­ tation with larvae. Biopsy 10 days after infestation showed that the concentration of DDT in body fat had increased from an earlier biopsy average of 1319  163 (S.E.) ppm to 3105  1071 ppm as a result of fat mobilization and a partial failure of metabolism and excretion to keep pace with the DDT so mobilized. The surviving rats had recov­ ered fully 38 days later; at that time the DDT concentra­ tion in their fat was only 874  10 ppm, because they had increased their fat stores into which the remaining DDT and that accumulated from continuing dietary intake were distributed. The reduction of food intake and the resulting loss of body weight produced by severe, nonfatal trichino­ sis were adequate to account for the initial increase in the concentration of DDT-derived material, and the subsequent recovery of weight was adequate to account for the final decrease in the concentration of this material stored in the fat of rats with gastrointestinal phase of this disease. Thus, all of the observed changes could be explained in terms of body weight, as was true for simple food deprivation (see this Section above). A different and less understood kind of interaction is that in which chemicals appear to reduce resistance to infec­ tion or to increase the virulence of an infecting organism.

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

Inasmuch as some compounds are antibiotic, it is logi­ cal that others may be probiotic. However, although some antibiotic reactions have been studied in great detail and are well understood, no careful study has been made of any probiotic interaction although it is well-known that low doses of antibiotics increase body weight gain in many species (Cababrese and Baldwin, 2001a). The anti­ biotic reactions show dosage-response relationships, and the compounds tend to be specific for groups of microor­ ganisms and sometimes for species or even strains. These same characteristics are likely to characterize any genu­ ine probiotic reactions. Examples of probiosis have been reported, but more investigation is needed to establish their validity and much more work is required to learn whether they are of practical importance and, if so, under what conditions. Although the implications of some individual reports of probiosis seem to have been exaggerated, the broadest theoretical implications have been neglected. In the absence of systematic study, it is impossible to exclude the possibility that probiosis is as important as antibiosis. This would certainly be true if chemical carcinogenesis was, in fact, a form of probiosis by alteration of a virus, a relationship strongly suggested by the work of Price et al. (1972), or, more generally, by influencing either a virus or its host, or both, as is now becoming evident for a human cancer caused by the hepatitis B virus (Beasley, 1988). Reports of probiosis involve polychlorinated biphe­ nyls and duck hepatitis virus in ducks (Friend and Trainer, 1970a); p,p’DDT or dieldrin and the same virus in ducks (Friend and Trainer, 1970b); and lead nitrate and Salmonella typhimurium in mice (Hemphill et al., 1971). Enhanced lethality of encephalomyocarditis (EMC) virus infection in suckling mice as a function of topical exposure to a combination of insecticides was reported by Crocker et al. (1974). Subsequent studies (Crocker et al., 1976) extended these observations to show that the responsible component was the insecticide carrier, the solvent emulsifier system in which the insecticides were prepared. On the contrary, Menna et al. (1980) exposed suckling CD1 outbred mice topically to insecticide carrier and found a decreased sensitivity to infection with lethal doses of influenza virus type A/PR8/34 (HONI) compared with untreated or mock-treated control mice. The decreased sensitivity was evidenced by significant increase in mean survival percentage of the mice after inoculation with infec­ tious agent. The decreased sensitivity was virus-dose-related and occurred within a dose range of 2–8 times the LD 50. One report (Wasserman et al., 1969) suggests that any change in response to infection may be complex but by necessity may involve the immune system. Rats given a 200-ppm aqueous suspension of DDT of unstated stabil­ ity as their only source of water for 35 days not only had heavier livers but slightly heavier adrenals and lighter spleens. The DDT-treated animals showed a rise of serum albumin and some globulin fractions but a decrease of

77

other globulin fractions. Although DDT alone caused a slight increase in the size of the adrenal, it tended to inhibit the greater increase produced by surgery. Whereas the average titer of antibodies to ovalbumin in rats receiving DDT was slightly less than that in controls injected with ovalbumin in the same way, the range of titers in differ­ ent animals in the same group was so great that the results were difficult to evaluate. It has not always been possible to confirm reports of a relationship between disease resistance and the intake of a chemical. For example, it has been reported that change in the phagocytic activity of white blood cells is an indication of early intoxication by DDT (Kun’ev, 1965). To test this report, DDT was given to male rats by stomach tube at a rate of 0.25 mg/kg/day for 31 days. Blood taken at intervals from 10 of these rats and from 10 controls was incubated with Staphylococcus epidermidis. The proportion of white cells ingesting bacteria and the average number of bacteria ingested per cell were measured. The same measurements were made at intervals on white blood cells from 15 dosed and 15 control rats after they had received bacteria by intracardiac injection. There was no statistical difference in phagocytic activity between the dosed and control rats in either the in vitro or the in vivo study (Kaliser, 1968). Temperature  The interaction of temperature and the effects of foreign chemicals is complex, but it must be taken into account in the design and interpretation of experiments. Such interaction is most likely to occur in connection with compounds that influence temperature control or metabolic rate, but is not confined to compounds known to have one of these actions. Disturbances and tem­ perature control are more likely to be important in small animals (such as rats and mice) or young animals, simply because their control of body temperature is imperfect at best. No matter what the size of the animals, the investi­ gator should record both the ambient temperature and the body temperature in any study involving temperature as a variable. In some instances, skin temperature or the tem­ perature of the extremities should be recorded because it may be critical but distinctly different from the body or visceral temperature. For most compounds, minimal toxicity occurs at some temperature between room temperature and thermal neu­ trality, that is, the temperature at which the animal con­ sumes least oxygen while at rest. In such instances, toxicity increases at temperatures both below and above this point, so that a graph of m ortality or other measure of toxicity against ambient temperature is U-shaped. Examples include ANTU (Meyer and Karel, 1948), parathion (Baetjer and Smith, 1956), warfarin, strychnine, and several com­ mon solvents (Keplinger et al., 1959). Apparently no com­ pound is known in which the opposite relationship exists, that is, maximal toxicity at some intermediate temperature with lesser toxicity at both lower and higher temperatures.

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A smaller number of compounds demonstrate a more or less continuous increase in toxicity corresponding to increasing ambient temperature. In this instance, the graph of toxicity against temperature may be thought of as the right-hand branch of a U-shaped curve. It is often an open question whether the remainder of the curve would be demonstrated if sufficiently lower temperatures were inves­ tigated. In any event, continuously increasing toxicity with increasing temperature has been found for dinitrophenol (Fuhrman et al., 1943; Keplinger et al., 1959) and picro­ toxin (Chen et al., 1943). The effects in mice of high and low environmental tem­ perature on the maternal and fetal toxicity of dinoseb and on the disposition of dinoseb were studied by Preache and Gibson (1975). Swiss-Webster female mice dosed with dinoseb were maintained at an increased environmental temperature (32°C) for 24 h or a decreased temperature (0–6°C) for 1.5–4 h. Increased temperature lowered the LD 50 for single injections of dinoseb and reduced temperature had no effect on the LD 50. Exposure of pregnant mice to the increased temperature increased maternal mortality, decreased fetal body weight, and increased the frequency of fetal anomalies relative to animals maintained at normal temperatures. Clearance of dinoseb from plasma or other tissues was not affected by exposure to high or low envi­ ronmental temperatures. Very few compounds may be more toxic at lower tem­ peratures so that the graph of toxicity against temperature may be considered the left-hand branch of a U-shaped curve. Whether higher temperatures would complete the curve is often unknown. According to Bogdanovic (1961), picrotoxin is an example but, as already noted, Chen et al., (1943) found the opposite in a very careful study. At least for some compounds that affect body tempera­ ture, there are critical ambient temperatures above which compounds cause a rise of body temperature and below which they cause a fall. Different compounds often have different critical temperatures in the same species (Shemano and Nickerson, 1958, 1959). If the change in body tempera­ ture is sufficient it may be a major cause of death. Even in the absence of a drug, an ambient temperature of 38°C is lethal to about 50% of mice in 3 h at a relative humidity of 20% (Adolph, 1947). Rats fed malathion at a dietary rate of 4000 ppm died sooner than controls when both were clipped and exposed to an ambient temperature of 1.5°C, but only after their body temperature had fallen to 18°C, which is half of normal body temperature (Marton et al., 1962). As Keplinger et al. (1959) pointed out, it may be difficult to decide whether the stress of heat or cold renders an ani­ mal more susceptible to a compound or whether the com­ pound renders the animal more susceptible to heat or cold. In fact, cold causes reactions of tetanus and hyperrespon­ siveness of the spinal cord similar in some respects to the reactions caused by strychnine and some other compounds (Koizumi, 1955; Brooks et al., 1955).

Hayes’ Handbook of Pesticide Toxicology

Many studies of the interaction of temperature and toxic­ ity are carried out in nonacclimated animals. This was true of most of the studies cited in this section. As shown in a paper by Johnson et al. (1963) and the associated discussion, accli­ matization may alter or even reverse the effect of either heat or cold. For example, Craig pointed out in the discussion that the toxicity of DFP, sarin, and atropine to rats and mice was increased by exposure to cold only if the animals were unac­ climatized. This means that the conditions of each study must be stated clearly. It does not mean that investigations of unac­ climatized animals are unimportant. People may encounter foreign chemicals at ambient temperatures to which they are not accustomed. The use of hypothermia in medicine is the most dramatic, but perhaps not the most important example. Temperature can affect absorption, distribution, and also action. As measured by excretion of paranitrophe­ nol, parathion is absorbed more rapidly from human skin at higher ambient temperatures (Funckes et al., 1963; Wolfe et al., 1970). The maximal average increase in absorption at 40.56°C compared with 14.44°C is apparently on the order of a factor of 4 but may be increased by a factor of 10 or more for the first few hours after exposure (Funckes et al., 1963). Some differences in action at different temperatures may be explained on the basis of dosage at the tissue level, as is true of chlorpromazine (Berti and Cima, 1954). However, this is not always true. For example, the cen­ tral nervous system depressant norpipanone is about three times more toxic to mice at 29°C than at 18°C ambient temperature (Herr et al., 1953). Although the difference is explained at least in part by the fact that, following identi­ cal doses, the concentration of the compound in the brain is about 40% greater at the higher temperature, there may be an inherent difference in the reactivity of the tissue. A higher dosage (96 mg/kg) and a resulting higher concen­ tration in the brain (30 ppm) were required to produce the same effect (LD 50) at 18°C than the dosage (33 mg/kg) and brain level (6.7 ppm) required at 29°C (Herr et al., 1954). Another example involves the action of DDT on the isolated frog heart. Hoffman and Lendle (1948) found that, in December at a low room temperature, a concentration of 300 ppm was required to produce the same effect that could be produced by 1 ppm or even 0.1 ppm in June at a temperature of at least 22°C. The same compound may produce qualitatively different effects at different temper­ atures. Fatal doses of chlorpromazine given to mice at an ambient temperature of 38°C cause violent convulsions, but at 13°C they cause prolonged central depression (Berti and Cima, 1955). The quantitative differences in toxicity associated with temperature are often small but are sometimes dramatic. Cold increased the toxicity of reserpine to unacclimatized mice by a factor of 1200 (LD 50, 0.0 15 mg/kg at 20°C, compared with 18.84 mg/kg at 30°C) (Johnson et al., 1963). Cold increased the susceptibility of rats to isoprenaline by factors of about 1,000 in males and 10,000 in females

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

(Balazs et al., 1962). Reviews of temperature effects include papers by Fuhrman and Fuhrman (1961), Ellis (1967), and Weihe (1973). A 2-year study of temperature effects in rats was performed by Yamauchi et al. (1981). In concluding this subsection, it is worthwhile to remind the reader that thermogenesis has been reasonable wellunderstood, being a result of uncoupling of oxidative phos­ phorylation potentially in all tissues (White, Hardler and Smith, 1973). Therefore, all uncouplers of oxidative pho­ phorylation will have a temperature effect and vice versa temperature will effect their toxicity. In addition to shivering thermogenesis, Himms-Hagen (1983) distinguished between non-shivering and diet-induced thermogenesis with the for­ mer primarily occurring in brown adipose tissue (HimmsHagen, 1985). 2,3,7,8-Tetra-chlorodibenzr-p-dioxin has been shown to reduce thermogenesis in this tissue (Weber et al., 1987). Probably many lipophilic pesticides affect this tissue but thus far little attention was paid to it. Pressure and Altitude  Pressure resulting from altitude may be a factor in the toxicity of any compound, especially one that influences cardiorespiratory function. An example involves the greater toxicities of red squill and digitalis at altitudes higher than most communities (Ward et al., 1940). Strychnine is also more toxic at high altitudes (Moore and Ward, 1935). On the other hand, a difference in pressure does not always create a difference in toxicity. Of practical interest is the finding that dichlorvos, at exposure levels far in excess of those proposed for the disinfection of aircraft, exhibited no toxicity to people at 2438 m, a cabin altitude seldom exceeded in normal airline operations of pressur­ ized aircraft (Smith et al., 1972). New problems of toxicology have arisen because sub­ marines and spacecraft may remain out of contact with the atmosphere of the earth for long periods of time. Maintenance of a small closed atmosphere offers a possi­ bility for the accumulation of various gases and vapors that dissipate rapidly in ordinary situations. The toxicity of any­ thing in the small space may be influenced by the fact that the pressure may not be that to which we are accustomed. Elaborate equipment for the study of these problems was first described by Thomas (1965). This issue is not trivial, because continuous exposure above a threshold is more toxic than any other exposure scenario. Changes in barometric pressure that occur in a labora­ tory as a part of changes in weather influence the activity of mice (Sprott, 1967) and rats (Olivereau, 1971); increased pressure and sudden falls were associated with increased spontaneous activity, whereas gradual decreases in pres­ sure had the opposite effect. It is well known from physics that all phenomena of nature are temperature-dependent unless the change in tem­ perature is so small that its effect is not measurable. The same can be said for pressure and volume except for a few limiting conditions. For the most part these variables play a

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minor role in toxicology for obvious reasons. Most labora­ tory animals are homeotherms, which means that biologi­ cal processes take place at almost constant temperature. Changes in ambient pressure are also comparatively small and very few species change their volume dramatically. However, under extreme conditions (hibernators, high alti­ tude, some frogs and toads) the dependence of c  t on these variables could possibly be established. Therefore, it should be noted that we are aware of the fact that the fun­ damental law of toxicology as formulated in this chapter should be written as (c  t  k)T,p,v or (c  t  k  W)T,p,v. Light and Other Radiation  Although this section is con­ cerned with the biological effects of light on toxicity, it should be recalled that some compounds undergo chemical change when exposed to radiation. Some of these changes have been demonstrated in pesticides and others may occur. Reactions in the upper atmosphere are considered important in degrading a variety of airborne compounds. Some pesticides known to be susceptible to photo­ dynamic action include p,p’-DDT (Roburn, 1963), p,p’DDE (Roburn, 1963), p,p’-DDD (Roburn, 1963), aldrin (Roburn, 1963), dieldrin (Robinson et al., 1966; Roburn, 1963). and endrin (Roburn, 1963). It has not been proved that these purely physicochemical changes are of any prac­ tical importance in the toxicity of any pesticide. There is some evidence that poisoning of crop workers may result from residues of paraoxon in fields treated with parathion (Milby et al., 1964). Conversion of parathion to paraoxon and other derivatives has been demonstrated in the labora­ tory (Frawley et al., 1958; Payton, 1953). However, it is not clear what factors favor the production and persistence of enough paraoxon in the field to produce poisoning. An old but still useful review of photodynamic action and diseases caused by light is that of Blum (1941). Weihe (1976) thoroughly reviewed the effect of light on labora­ tory animals. The need for standardized lighting in animal rooms is well documented (Kaitz and Auerbach, 1979; Reiter, 1973; Robinson and Kuwabara, 1976; Weisse et al., 1974). Ionizing Radiation   The biological effects produced by X-rays and other ionizing radiation have been studied extensively. A description of these effects is beyond the scope of this volume, even though gamma rays from radio­ active cobalt were used to sterilize screw-worm flies to eradicate this destructive species in the southeastern United States. A useful review of the biological effects of ionizing radiation is that of Schwan and Piersol (1954, 1955). Ultraviolet Radiation  In addition to the direct photo­ chemical action mentioned at the beginning of this sec­ tion, ultraviolet radiation with a wavelength in the range of 0.29–0.32 m is responsible for sunburn. Ultraviolet light

80

also produces “farmer’s skin,” an increased incidence of skin cancer, and the conversion of 7-dehydrocholesterol or a similar precursor in skin to vitamin D. These effects and the general action of sunlight were reviewed in detail by Blum (1945). Visible Radiation  Difference in the intensity and wave­ length of light within the visible range may influence the production of a variety of physiological effects which may interact with the effects of toxic substances or even become manifest only in the presence of such substances. Kueter and Ott (1964) reported acceleration of the appear­ ance of carcinoma, increased aggressiveness, and reversal of the sex ratio as effects of artificial light from various commonly used sources. Animals in toxicology laboratories are routinely housed in cage racks with multiple shelves to maximize the use of animal room space. Greenman et al. (1982) reported the influence of shelf level on retinal atrophy in mice. It had been shown earlier (Noell et al., 1966) that 24-h exposure of unrestrained rats to light from ordinary fluorescent bulbs causes irreversible damage. Greenman et al. (1984) evaluated the association between cage shelf level and spontaneous and induced neo­ plasms in about 21,000 mice being used in a study to evalu­ ate the carcinogenicity of 2-acetylaminofluorene. There was evidence for a shelf level influence on five of the six major spontaneous neoplasms noted. Time to onset of uterine pol­ yps and reticular cell sarcomas was significantly delayed on the top shelf of five of six animal rooms. Also, there was significant delay in the onset of lymphomas, adrenocortico­ adenomas, and lung alveolar cell tumors on the top shelf when data were combined from all six animal rooms, but these delays on the top shelf were significant in no more than two of the six animal rooms when rooms were ana­ lyzed separately. Thus, this study provided evidence that a shelf level must be considered in the design and analysis of carcinogenesis studies. In contrast to these observations, Haseman (1988) reported a lack of cage effect on liver tumor incidence in B6C3F1 mice constituting a total of 89 dosed groups showing increased liver tumor incidence. Experimental design protocols that include random assignment of columns of cages to dosed and control ani­ mals, periodic rotation of cage location, and indi­vidual caging of animals reduced the likelihood that differences in lighting or other factors associated with cage placement of animals could influence the results of toxicity studies. Photoperiodicity  Photoperiodicity of visible light deter­ mines or synchronizes circadian rhythms and, in com­ bination with changes in temperature, is responsible for seasonal changes in physiology (see Section 1.5.7.1). Photosensitization  Some chemicals make cells more sus­ ceptible to the action of light, especially ultraviolet light.

Hayes’ Handbook of Pesticide Toxicology

Effects have been reported to result from wavelengths rang­ ing from 0.29 to 0.50 m (Daniels, 1965). Most compounds with this property are fluorescent (Blum, 1941), Although photosensitization usually affects the skin of vertebrates, other tissues are not immune. For example, the perfused turtle heart was arrested by a porphyrin preparation when exposed to light, but not in the dark. The reaction was not caused by a diffusible toxin. A second heart perfused in the dark with perfusate from the first heart was not affected (Rask and Howell, 1928). Actions on other vertebrate tis­ sues as well as free-living cells, viruses, and proteins including enzymes have been demonstrated (Blum, 1941). One of the outstanding characteristics of photodynamic processes is that they occur only in the presence of molecu­ lar oxygen. However, photodynamic uptake of oxygen dif­ fers strikingly from normal oxidative metabolism in regard to respiratory quotient and sensitivity to heat and chemical inhibitors (Blum, 1941). Whereas chemical photosensitization generally is acti­ vated by ultraviolet light, visible light may be an activator also, at least in some organisms. For example, although paramecia are not injured or sensitized to heat by visible light of high intensity, they readily are killed by this light in the presence of photodynamic dyes, and they are sensitized to heat by sublethal dosages of light. Cells so sensitized are killed when subjected to otherwise harmless temperatures. If the light and heat are applied in the reverse order, no ill effects are observed (Giese and Crossman, 1946). In most instances, the biochemical basis of photosensi­ tization is not understood, but it certainly can involve basic components of protoplasm. Deoxyribonucleic acid suspen­ sions become less viscous when irradiated in vitro in the presence of eosin, methylene blue, 1,2-benzanthracene or 20-methylcholanthrene. It is thought that the reaction involves depolymerization (Koffler and Markert, 1951). Photosensitization has been caused in one species or another by a wide range of compounds. In addition to the porphyrins, the following materials have caused some degree of photosensitization in humans: methylene blue, many phe­ nothiazine compounds, many furocoumarin compounds, anthracene and acridine derivatives, 5-methoxypsoralen (the active principal of oil of bergamot used in perfumes) and related materials from other plants, griseofulvin, demethylchlorotetracycline and some other antibiotics, sulfonamides and their derivatives (including some oral hypoglycemic agents), bithionol, hexachlorophene, and other miscellaneous drugs. Note that many of these com­ pounds consist of three aromatic or heterocyclic rings in a linear configuration. Substitution with sulfur or nitrogen may lead to an increase in photosensitizing capacity (Daniels, 1965). A number of pesticides have chemical structures sug­ gesting they might act as photosensitizers. This property has been observed in connection with oxythioquinox, phe­ nothiazine and griseofulvin.

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

Porphyrins are probably the cause of more frequent and more serious photosensitization in humans than all other materials combined. However, photosensitization does not occur in all cases in which the concentration of porphy­ rins in the blood and excreta is increased. It is important to note that the chemical most effective for disturbing this metabolism is a now banned pesticide, hexachlorobenzene. Detailed reviews have been written by DeMatteis (1967), Schmid (1966), and Jaffe (1968), and Courtney (1979). Circadian and Other Rhythms  A wide range of biologi­ cal rhythms and their bases have been reviewed at length (Sollberger, 1965). The word “circadian,” from the Latin circa (about) and dies (a day), refers to the rhythmic rep­ etition of certain phenomena in living organisms at about the same time each day. Of course, some circadian rhythms have been com­ mon knowledge for centuries. Some animals are nocturnal and others are diurnal. One might suppose that this differ­ ence in activity pattern depended merely on species differ­ ences regarding direct response to light. Experimentation has revealed that the situation is often not so simple. The activity pattern may persist, with or without modification, when an animal is placed in continuous darkness or contin­ uous light. Even when experimentally imposed conditions of light change the pattern, the regulation may be tempo­ rary and the natural rhythm may eventually exert itself. For example, although Siegel (1961) found that the diur­ nal feeding pattern of rats disappeared in 6–10 days in rats transferred to continuous light, Wiepkema (1966) found that mice reared for two generations in continuous light showed a marked circadian rhythm in their feeding. The persistence of circadian rhythm in the absence of any known external clue to the passage of time is one kind of evidence that has led to the hypothesis of the “physio­ logical clock,” Although the anatomical location and mode of action of such a clock is not known precisely, there is convincing evidence that some circadian rhythms are endogenous. Both endogenous and exogenous circadian rhythms are adjusted and regulated by photoperiod. Under natural conditions, photoperiodicity depends on the movement of the earth. Rotation of the earth produces succession of day and night. Revolution of the earth around the sun produces the seasons. In the temperate and arctic zones, the days are longer in summer and shorter in win­ ter. Furthermore, the difference in the length of daylight at these seasons increases progressively from the equator to the poles. Consequently, during spring and fall, the rate of change in the relative duration of light and darkness is greatest at the poles and least near the equator. The ability of organisms to use photoperiodic cycles as clues to impending seasonal change implies that they possess the ability to distinguish between long and short lengths of daylight. This ability is one factor in the complex adapta­ tion of organisms to their environment.

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The complicated and varied effects of photoperiodic­ ity on organisms have been abundantly demonstrated by experiments designed for that purpose. There is a danger that the possible importance of photoperiodicity will be forgotten in experiments designed for other reasons. Many modern animal rooms have only artificial lighting, but the lighting cycle and the adaptation of the animals to it fre­ quently are not mentioned in descriptions of methods used in toxicological studies. It appears that the effects of circadian rhythms and pho­ toperiodicity in invertebrates, especially insects, have been studied more thoroughly than those effects in mammals. Valuable reviews of the physiology and ecology of photo­ periodism in insects have been written by Beck (1963) and Danilevskii (1965). It is impossible to go into the matter in detail here. It is pertinent to record that one or more spe­ cies of insects or mites show definite diurnal variation in their susceptibility to some pesticides, including dichlorvos (Polcik et al., 1964), methyl parathion (Cole and Adkisson, 1964), DDT (Beck, 1963), and potassium cyanide (Beck, 1963). Probably some of the information on invertebrates would be of value in connection with studies in mammals. However, it is already clear that circadian rhythms are important in a number of physiological functions of mam­ mals, including their susceptibility to some poisons. Circadian Rhythms in Mammals  In mammals, as in insects, endocrine functions may be influenced directly or otherwise by light and may involve circadian rhythms. For example, hepatic tryptophan pyrrolase and its circulating substrate, whole-blood tryptophan, have a circadian rhythm in mice that is practically eliminated by adrenalectomy (Rapoport et al., 1966). However not all liver enzymes are so greatly influenced by adrenalectomy. Civen et al. (1967) showed that the rhythmicity of tyrosine ketoglutarate trans­ aminase (TKT) is little altered after adrenalectomy. The same authors Civen et al. (1967) noted that TKT is rapidly induced by various agents but that phenylalanine pyruvate transaminase (PPT) is not induced during the same time period and does not show circadian variation. On the basis of this and some other evidence, they sug­ gested that the sensitivity of an enzyme’s regulating system to inducing agents may be related to the inherent circadian rhythm of the enzyme. The exact function, if any, of the pineal gland (epiphy­ sis) is still in doubt. Because of its histology and the nature of its embryonic origin, it has been suspected for a long time that the structure has an endocrine function. This pos­ sibility, with special reference to neurohormonal control, seems to gain support from the demonstration (Quay and Halevy, 1962) that the pineal gland is rich in serotonin. Studies of the gland illustrate the complex interrelation that circadian rhythms may show in one small detail of mammalian physiology. In the rat, the serotonin content of

82

the gland shows a circadian rhythm (Quay, 1963; Snyder et al., 1965) which is somewhat modified by the estrus cycle (Quay, 1963). The rhythm persists in rats kept in the dark or in rats whose eyes have been removed, pro­ vided the animals are otherwise intact. The rhythm is abol­ ished in intact rats by continuous light and also abolished by interruption of sympathetic innervation (Fiske, 1964; Snyder et al., 1965). The rhythm is changed in a matter of hours by change in photoperiods (Quay, 1963). The rhythm is not affected by removal of the pituitary, thyroid, adre­ nals, or ovaries (Snyder et al., 1965). The rat pineal gland also shows circadian rhythms for hydroxyindole O-meth­ yltransferase (HIOMT) (Axelrod et al., 1965) and endog­ enous melatonin (Quay, 1964). However, these rhythms are opposite in phase to that of serotonin and also differ in that they do not persist in animals kept in the dark. The rhythms for HIOMT and melatonin are directly responsive to light. All three rhythms (HIOMT, melatonin, and sero­ tonin) are interrupted by removal of the superior cervical ganglia (Fiske, 1964; Snyder et al., 1965; Wurtman et al., 1964). Because the nerve pathways are probably noradren­ ergic, McGeer and McGeer (1966) explored the possibil­ ity that there might be a circadian rhythm in the ability of the nerve endings of the pineal gland to form noradrena­ lin. They found such a rhythm in the activity of tyrosine hydroxylase, the rate-controlling enzyme in the synthesis of noradrenalin. Some circadian rhythms involve reactions known to be of fundamental physiological importance. For example, Spoor and Jackson (1966) showed that the beating rate of rat atria decreased more in response to a standard concen­ tration of acetylcholine if they were isolated at 1100 hours than if isolated at 2300 hours. The food intake of rats nor­ mally follows a circadian rhythm, the details and modifica­ tion of which were studied by Siegel (1961). The preceding examples involve rhythms with a single major peak and a single major trough during each 24-h day. The rhythm is either a direct response to the periodicity of light or, if endogenous, at least is synchronized by light. Lindsay and Kullman (1966) reported that the survival time of female mice given a standard dose of sodium pentobarbi­ tal varied during a 12-h period in such a way that the graph showed several inflections. Although this result is unex­ plained and unconfirmed, it is not unique. A similar result was reported for the susceptibility of boll weevils to methyl parathion (Cole and Adkisson, 1964). The weevils showed greatest resistance at the beginning of the light period no matter whether it started at 0600 hours (14-h photophase), 0700 hours (12-h photophase), or 0900 hours (10-h photo­ phase). Regardless of the length of the photophase, peaks of resistance recurred at intervals of about 6 h with inter­ vening troughs of susceptibility. Why the 3-h cycle cor­ responded with the sampling interval was unexplained. In any event, it is interesting that both examples of multiple

Hayes’ Handbook of Pesticide Toxicology

peaks involve susceptibility to a foreign compound. Nothing is known of the enzymatic or other physiological basis for the reported phenomenon. Typical circadian rhythms (one peak and one trough during a 24-h day) are involved in the responses of several mammals to a number of foreign chemicals. Such cycles of susceptibility were observed at least as early as 1949 (Carlsson and Serin, 1950). As reviewed by Sollberger (1965), a 24-h rhythm in sensitivity of mammals to a number of drugs has been reported. Compounds involved include insulin, hormones, narcotics, sedatives, tranquil­ izers, bacterial toxins, and carcinogens. Other examples include lidocaine (Lutsch and Morris, 1967), methopyra­ pone (Ertel et al., 1964), nikethamide (Carlsson and Serin, 1950), and pentobarbital (Davis, 1962). Human circadian rhythms can persist in continu­ ous darkness; social cues are sufficient to entrain them (Aschoff et al., 1971). It is clear that circadian rhythms or the effects of light periodicity should be considered when there are un­explained differences in the results of different laboratories or in the results of the same laboratory at different times. Examples of the influence of pesticides on the cir­ cadian rhythms or the effect of circadian rhythms on the toxicity of pesticides are not common, but Nicolau (1982) reported on the effects of pesticides on the circadian time regarding the structure of the thyroid, adrenal, and the tes­ tis in rats. Four herbicides, a fungicide, and two insecti­ cides were tested. A wide variety of rhythm alterations was found. There was significant desynchronization of thyroid and adrenal gland functions. In contrast, there was almost no effect on the rhythms in the testis. Nicolau (1983) also reported on the effect of dichlorvos and trichlorfon on cir­ cadian rhythms of RNA, DNA, and protein synthesis in the rat thyroid, adrenal, and testis during exposure to these chemicals for 90 days. Prolonged exposure of Wistar rats to trichlorfon at a concentration of 10 ppm in the diet led to marked changes in the circadian rhythm of the thyroid DNA and protein and adrenal DNA content, phase altera­ tions in thyroid RNA and adrenal DNA rhythms, and marked decrease in the amplitude of the adrenal DNA and protein rhythms. Exposure to dichiorvos in the diet at a concentration of 5 ppm led to phase alterations, with­ out a change in the time-qualified mean, of the circadian rhythms in DNA, RNA, and protein content and marked decrease in amplitude of the DNA rhythm in the thyroid and adrenal. There were no alterations in the rhythms of testicular function with either of the pesticides studied. Other Factors Influencing Toxicity  Undoubtedly, many factors in addition to those discussed in the preceding sec­ tions may influence toxicity under certain circumstances. However, these other factors are probably not of major importance in mammalian toxicology.

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

Seasonal Differences  Seasonal differences are of tremen­ dous importance in the physiology of cold-blooded animals and in their responses to toxicants. Presumably, similar dif­ ferences would hold for mammals that hibernate, but the question has received little attention. Seasonal variation in the LD 50 of mice treated with organic solvents has been reported (Wolff, 1985), but Gaines and Linder (1986) found no seasonal pattern for either parathion or DDT although the bimonthly saw-toothed pattern for DDT was the same for male and female rats. Relative Humidity  Presumably, relative humidity might influence the reaction of an animal to a toxin in any situation in which humidity was already critical for the animal’s health, for example, maintenance of normal body temperature in a hot environment. Such an interaction would seem most likely in connection with compounds that increase heat production or influence temperature control. However, no instance of such an effect on toxicity seems to have been reported. What has been reported is an effect of relative humidity on the absorption of insecticides and, therefore, on their availability for evaporation or absorption from surfaces. It has been shown that parathion is absorbed more rapidly by the human skin at higher temperatures (Funckes et al., 1963; Wolfe et al., 1970). What part humidity (especially from sweat) may play in the process apparently is not known. It is probable that mammals may be able to absorb pesticides from their skin surfaces more readily if the sur­ face is moist, because absorption is usually increased after application of occlusive patches. Aquatic Factors  Because it is sometimes suggested that fish or other aquatic organisms be used for bioassay of toxicants that might influence mammals, it is necessary to record that the welfare of aquatic organisms is influenced by several environmental factors that have little meaning for land ani­ mals. Most important is continuous exposure in an aquarium. pH  The influence of pH on toxicity often is explained eas­ ily in terms of the availability of toxicant. Toxic ions or alkaloids may be much more soluble or easily absorbable at one pH than at another. Water  Hardness Henderson and Pickering (1957) found that water hardness had a significant effect on the toxicity of trichlorfon to fathead minnows but no significant effect on the toxicity of nine other organic phosphorus com­ pounds they studied. Water hardness had little or no effect on the toxicity of chlorinated hydrocarbon insecticides to fish (Henderson et al., 1959). Chlorine Content Ordinary  tap water may contain enough free chlorine to kill some fish. This must be kept in mind in bioassays on fish.

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It must be pointed out that aquatic toxicity is kinetically similar to continuous inhalation exposure. Therefore, tox­ icity determined in an aquarium in a species that spends all its time in water will strictly obey Haber’s rule, unless the experiment harbors some uncontrolled variables discussed by Hayes in preceding subsections.

1.5.7.2  Species Differences Due to Kinetics Kinetics (K) is the mathematical description of the time course of a chemical in an organism as affected by absorp­ tion, distribution, biotransformation, and excretion. It should be emphasized that species differences in K may be due to any of these processes. A widely held reductionist view­ point led to only biotransformation being given due scrutiny as a potential cause of species differences in the disposition of xenobiotics (Caldwell, 1982). However, an astute inves­ tigator ought to embark upon studying the disposition of a new compound without such bias. The following discussion will demonstrate the importance of bio-transformation for species differences, but it will also point out the pitfalls of failing to take into consideration other important processes involved in the disposition of xenobiotics. Absorption Gastrointestinal Absorption  In general, ­ gastrointestinal absorption of xenobiotics was thought to be similar between species. The work of Dreyfuss and colleagues illustrates the fallacy of this assumption (Dreyfuss et al., 1978). Absorption of nadolol [calculated from AUC after intraperi­ toneal (ip), intravenous (iv), and oral (po) dosing] was essen­ tially complete in the dog, substantially less in humans, and quite limited in the rat (Table 1.11). Urinary and fecal excre­ tion of nadolol support the bioavailability data. However, excretory data further indicate that, in addition to the non­ absorbed portion of this compound, biliary and possibly nonbiliary sources also contribute to the fecal excretion of

Table 1.11  Absorption and Excretion of Radioactivity in Rats, Dogs and Humans after Nadolol Dosagesa Species

Dose Route (mg/kg)

Percent of dose excreted Urine

Feces

Percentage of dose absorbed

Rat

20 20

po ip

11 62

84 31

18 (100)

Dog

25 25

po ip

76 75

28 12

102 (100)

Human

2 2

po ip

25 73

77 23

34 (100)

a

Modified from Dreyfuss et al. (1978).

Hayes’ Handbook of Pesticide Toxicology

84

this compound. Calabrese (1984) reported evidence for spe­ cies differences in the absorption of at least 38 compounds, indicating that nadolol may not be an exceptional case. It is more likely that the possibility of species differences due to differences in absorption has seldom been considered or examined. This is surprising because numerous characteris­ tics of the gut suggest that species differences in the absorp­ tion of xenobiotics may be expected. The rate-limiting barrier in the absorption of most xenobiotics is the unstirred water layer along the intestinal mucosa (Hayton, 1980). The effect of the unstirred water layer as a possible cause of species differences in absorp­ tion of xenobiotics has not been investigated. However, Thomson et al., (1983) studied the effect of the unstirred water layer on the absorption of fatty acids and cholesterol. These authors concluded that the thickness of the unstirred water layer may contribute to species differences in the absorption of lipophilic compounds, but other tissue-specific differences must also exist because species differences persisted when the unstirred water layer was diminished by stirring. Based on these considerations, it is reasonable to assume that the permeability of the gut for xenobiotics transported by passive diffusion can be species-dependent. Anatomical (allometric) considerations are another likely reason for species differences in intestinal absorp­ tion. The relative length of intestinal segments is quite vari­ able (Iatropoulos et al., 1986), and substantial functional differences exist between such species as ruminants and omnivores (Smith, 1986). Because most xenobiotics are transported across the gastrointestinal mucosa by passive diffusion, and because this transport is surface-area- and site-dependent, it can be expected that these factors will be responsible for species differences in some in-stances. Many xenobiotics are weak organic acids or bases. For such compounds, gastrointestinal absorption is dependent on the pH along the gastrointestinal tract. Table 1.12, modified from the work of Smith (1986), shows that each segment of the gut reveals considerable species specificity, with differ­ ences of up to two pH units. This translates into two orders of magnitude difference in terms of the concentration of the

Table 1.12  pH of the Gastrointestinal Contents of Various Speciesa Species

pH Stomach Jejunum Cecum

Colon

Feces

Monkey

2.8

6.0

5.0

5.1

5.5

Dog

3.4

6.6

6.4

6.5

6.2

Rat

3.8

6.8

6.8

6.6

6.9

Rabbit

1.9

7.5

6.6

7.2

7.2

a

Modified from Smith (1965).

undissociated versus dissociated moiety of a weak organic acid or base. Obvious consequences of such differences for absorption have been discussed by Shanker (1962). An additional factor that may result in species-dependent absorption of xenobiotics is the gastrointestinal flora. In gen­ eral, the microflora of animals is remarkably similar, although qualitative and quantitative differences have been reported (Smith, 1965). Notable deviations to this generalization do exist, such as the rabbit and humans (Table 1.13). In con­ trast to other species, the microflora in these two species is essentially absent in the upper gastrointestinal tract. Because absorption of some xenobiotics requires prior bacterial hydro­ lysis, some species differences may be due to differences in microflora. The example of cycasin is discussed by Rozman and Iatropoulos (1989). Cycasin is poorly absorbed by gno­ tobiotic animals; however, the aglycon of cycasin is readily absorbed. Therefore, species with bacterial -glucosidase activity in the upper small intestine readily absorb the aglycon (methylazoxymethanol), but species like humans, with very low levels of microflora in the upper gastrointestinal tract, may not absorb this compound to any major extent. Dermal Absorption  Species differences related to der­ mal absorption of xenobiotics have been more appreci­ ated (Calabrese, 1984). Dermal absorption of endogenous or exogenous compounds may vary by orders of magnitude (Kao et al., 1985). According to Dugard (1983), two fac­ tors are important in dermal absorption of chemicals: the appendages (sweat ducts, pilosebaceous ducts) in the early phase of absorption and the stratum corneum in the late and dominating phase of absorption. Both factors are highly species-dependent. Because the stratum corneum is much thicker in humans than in animals, human skin is usually less ­permeable for xenobiotics than is animal skin. However, the thinner stratum comeum in animals is often compensated for by a relatively thick hair cover, diminish­ ing direct contact of the skin with a xenobiotic. Sweat and pilosebaceous ducts also reveal great species variability.

Table 1.13  Number of Microbes and Their Distribution along the Gastrointestinal Tract of Various Speciesa Species

Stomach

Jejunum

Colon

Feces

Monkey

23

24

41

38

Dog

19

20

40

43

Rat

18

23

37

38

Rabbit

4

5

13

13

Human

2

4

10

—

a

Modified from Smith (1965) and Hallikainen and Salminen (1986). Expressed as log10 of viable counts.

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

Eccrine sweat glands are located in the pads of the extrem­ ities of all mammals. However, the general body surface of humans contains 100–600/m2 of coiled tubular sweat glands, whereas rodents and rabbits have none (Szabo, 1962). The number of pilosebaceous ducts in humans and pigs is similar (about 40/cm2), but rodents may have 100 times more (Calabrese, 1984). Moreover, biotransforma­ tions in skin that facilitate absorption also display great species variability (Kao et al., 1985). Another important potential rate-limiting step in the dermal absorption of chemicals is the cutaneous blood flow. Due to an important thermoregulatory function of the skin in humans as opposed to furred animals, there is a much more extensive vasculature in humans than in most mam­ mals (Calabrese, 1984). This brief discussion illustrates that species differences in the disposition of xenobiotics after dermal exposure may be due to numerous anatomi­ cal, physiological, and biochemical factors. Distribution  This process in the disposition of xenobiotics has rarely been considered as a potential cause of inter-spe­ cies variability. However, a closer scrutiny of the literature indicates that this may be an unjustified assumption. Plasma Protein Binding The disposition of clofibrate (clo­ fibric acid) is a case in point. Plasma protein binding of clo­ fibric acid reveals considerable species differences between mice, rats, and humans, which roughly correlates with the half-lives of this compound in these species (Table 1.14). Because clofibric acid is primarily eliminated in all three species by glomerular filtration without tubular reabsorp­ tion (pK  3), differences in the free fraction of this com­ pound in plasma of various species are likely to contribute greatly to the observed species differences. The other major factor is renal clearance (blood flow dependent). Additional factors that influence plasma protein bind­ ing may also be responsible for species differences, as dis­ cussed by Wilkinson (1983). Most important are species differences in the concentration of albumin, binding affin­ ity, and competitive binding of endogenous substances. Tissue Binding This is an area where information is scarce. Kurz and Fichtl (1983) reported good correlation

for the binding of drugs to muscle of man and rabbit. However, a more typical example is reported by Batra et al. (1986), which shows that interspecies variations in the tissue distribution of mitoxantrone are unpredictable and may vary by more than an order of magnitude. One frequently overlooked cause of species differences in the distribution of xenobiotics with large tissue accumu­ lation tendency (e.g., storage in fat) is the different rate of growth of mammals (Scheufler and Rozman, 1984). As Freeman et al. (1989) demonstrated using a physiologi­ cally based pharmacokinetic model, tissue and whole body growth contribute more to the distribution profile of hexa­ chlorobenzene than does excretion. Biotransformation  This is the best-documented cause for species differences in the disposition of xenobiotics (Caldwell, 1981, 1982). Very informative in this context is Walker’s (1980) compilation of monooxygenase activities in 65 verte­ brate species. The presence of cytochrome P450 and its asso­ ciated electron transfer components across broad taxonomic classes suggests that this system has arisen from some ancient genome. The most likely explanation for the vast species dif­ ferences in the expression of this genome is the evolutionary need to respond to changing diet, life style, and habitat. Phase I Biotransformations Caldwell (1981) illustrates the consequences of species differences in phase I bio­ transformation for the disposition of a number of amphet­ amines. Deamination is the major pathway of amphetamine biotransformation in rabbits and guinea pigs, whereas aromatic hydroxylation is the predominant route of bio­ transformation in the rat. The rhesus monkey utilized both pathways to a similar extent, and the marmoset neither one. Correspondingly, the marmoset excreted an administered dose unchanged, whereas the other species eliminated little of the parent compound, but rather the respective metabo­ lites. The broad tissue (liver and intestine) substrate speci­ ficity of monooxygenase isozymes is shown in Table 1.15. This table also illustrates the evolutionary importance

Table 1.15  Species Differences in Substrate Specificity of Monooxygenases in the Liver and Intestinea Species

Table 1.14  Plasma Protein Binding and Half-Life of Clofibric Acid in the Mouse, Rat, and Humana Species

Plasma protein binding (%)

Half-life (h)

Man

97

21

Rat

75

6

Mouse

45

2

a

Modified from Cayen (1980).

85

a

Benzo[a]pyrene hydroxylaseb

Ethylmorphine N-demethylaseb

Liver

Intestine

Liver

Intestine

Rat

0.33

0.14

3.8

ND

Mouse

0.15

0.10

3.4

ND

Rabbit

0.06

0.84

1.3

11.2

Guinea pig

0.07

0.37

1.4

8.8

Modified from Gregus et al. (1983) and Laitinen and Watkins (1986). Expressed as nmol/min/mg protein; ND, not detectable.

b

Hayes’ Handbook of Pesticide Toxicology

86

Table 1.16  Effect of Diet on Phase I Biotransformations in the Guta

Table 1.17  Urinary Excretion of Phase II Biotransformation Products of Organic Acidsa

Dietary condition Fat deficiency

Phase I biotransformation Decreased

Species

Fat excess

Decreased

Cholesterol excess

Increased

Cholesterol deficiency

Decreased

Copper excess

Increased

Selenium deficiency Cabbage

Naphthylacetic acidb Glucuronides Amino acid conjugates

Hydratropic acidb Glucuronides Amino acid conjugates

Rat

23

51

0

64

Ferret

72

19

91

5

Decreased

Rhesus monkey

3

83

0

75

Increased

a

Modified from Symchowicz et al. (1967). Expressed as percentage of dose; ND, not determined.

b

Brussels sprouts

Increased

a

Modified from Laitinen and Watkins (1986).

of diet in the development of monooxygenase activities, because rats and mice, rabbits and guinea pigs are similar, respectively, but omnivores differ greatly from herbivores. Composition of the diet may also play an important role in the biotransformation of xenobiotics on a day-to-day basis. Table 1.16 demonstrates that certain phase I biotransforma­ tions may be increased or decreased depending on dietary factors. It is thought that dietary components induce or inhibit monooxygenases and thereby alter phase I biotrans­ formation of xenobiotics with corresponding consequences of increased or reduced rates of elimination. Phase II Biotransformations  In the disposition of xenobiot­ ics, phase II biotransformations are usually more important than phase I reactions, although phase I reactions may be a prerequisite for subsequent phase II biotransformations in some instances. For example, insertion of a hydroxyl group or epoxidation (phase I biotransformations) does not change the lipophilicity (and hence reabsorbability) of a xenobiotic to any great extent. However, phase II biotransformations (gluc­ uronidation, sulfation, glutathione conjugation) increase water solubility very substantially, and hence increase the renal or biliary excretability of xenobiotic metabolites. The same factors affect phase II biotransformations that affect phase I biotransformations. Therefore, it is not surprising that vast species differences exist in the phase II enzyme-­dependent disposition of even structurally highly related xenobiotics (Table 1.17). Thus, accurate species-to-species predictions regarding phase II biotransformation-dependent disposition of xenobiotics remain elusive, which may or more often may not hamper the predictive value of one species for another in terms of toxicity. The crucial role of genetics in enzyme activ­ ities is clearly illustrated by Table 1.18, showing that phyloge­ netic relationships allow at least some generalizations. Excretion  This is the final and irreversible step in the dis­ position of xenobiotics. Consequently, any of the previous

Table 1.18  Predictive Pattern of Animal Biotransformation Reactions for Humansa Species

Prediction Good or fairb

Poor or invalidb

Rat

41

59

Other nonprimate

59

41

Rhesus monkey

92

8

a

Modified from Caldwell (1981). Expressed in terms of % of occasions predictable for humans in each category.

b

steps (absorption, distribution, biotransformation), as well as differences in excretion itself, may be responsible for species differences in the elimination of xenobiotics. In a simple case (e.g., inulin), when a compound is injected intravenously (no absorption) and does not bind to plasma protein, or does not distribute to tissues, or does not get biotransformed, and its only route of elimination is glomerular filtration, then the cause of species differences can be attributed solly to the rate of blood flow to the kidneys. However, in most instances the situation is much more complicated, as discussed previously for the individual steps in the disposition of xenobiotics. A few important examples follow that will illustrate the major factors determining excretion of xenobiotics. Urinary Versus Biliary Excretion  This point is best exemplified by the disposition of griseofulvin in rats and rabbits (Table 1.19). Rabbits excrete most of a dose of gris­ eofulvin as 6-demethylgriseofulvin in urine. This is to be expected, because the molecular weight of this compound is only 328. According to Hirom et al. (1976), molecules with molecular weight (MW)  350 tend to be preferen­ tially excreted in urine, whereas those between 350 and 700 are predominately excreted in bile. Because the molecular

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

weight of griseofulvin conjugates is about 500, it is not surprising that rats, which biotransform griseofulvin exten­ sively (phase II), excrete much of a dose in bile. This is an example of biotransformation being the critical step in the disposition of a xenobiotic. It is important to empha­ size that alternative possibilities ought to be considered in any given instance to ensure that species differences are resolved in the disposition of xenobiotics. This point was illustrated by the work of Migdalof and colleagues, using captopril (Migdalof et al., 1984). This weak organic acid is predominately excreted in urine as the parent compound by both dogs and monkeys. It has negligible plasma protein binding and biliary excretion. Yet, urinary clearance of captopril is about three times as rapid in monkeys as in dogs. The authors resolved this species difference by deter­ mining that active tubular secretion of captopril is about three times higher in monkeys than in dogs. Urinary Versus Fecal Excretion Often the elimination of a compound occurs by different routes in different spe­ cies, as shown in the case of indomethacin in the dog and the rhesus monkey (Table 1.20). Dogs excrete most of a dose in feces, whereas monkeys excrete the majority of a dose in urine. Both species excrete similarly large quanti­ ties of a dose in bile. Because dogs excrete most of a dose in bile as conjugates (MW  500), it is to be expected that

Table 1.19  Urinary and Biliary Excretion of Griseofulvin and/or Metabolites in Rats and Rabbitsa Ratsb

Rabbitsb

87

these hydrophilic indomethacin derivatives will not be reabsorbed unless they are hydrolyzed by intestinal bacteria to the reabsorbable parent compound or to phase I metabo­ lites (which have good bioavailability). Based on available experimental data, it is not possible to decide with cer­ tainty whether or not this is occurring in the dog. It appears that indomethacin undergoes enterohepatic circulation with repeated conjugation in the liver and deconjugation in the small intestine, with a gradual “loss” of conjugates into the large intestine. However, because almost all of fecal excre­ tion consists of indomethacin it is apparent that the large intestinal flora hydrolyzes the indomethacin conjugates. Limited reabsorption of indomethacin is not surprising (pK  4.5, colon pH  8), because more than 99.7% of indomethacin is ionized in the large intestine, which has a small surface area (compared to the small intestine). This does not allow for a sufficiently rapid shift in the mass bal­ ance to result in substantial reabsorption. The monkey also reveals extensive enterohepatic recy­ cling (57.7% of dose excreted in bile within 2 h). However, most of the biliary excretion consists of parent compound, which is readily reabsorbed. Furthermore, biliary conju­ gates appear to be hydrolyzed by small intestinal bacte­ ria followed by reabsorption, because “loss” into feces is comparatively small (about 10% of dose). In contrast to the dog, monkeys excrete most of a dose as phase I metabo­ lites (24.2% of dose) and indomethacin (10.5% of dose). Because indomethacin has a molecular weight of 358 and phase I metabolites have molecular weights of 220–345, these compounds are readily excreted in urine, as expected according to the work of Hirom and coworkers (Hirom et al., 1976).

Urine

Bile

Urine

Bile

Total

12

77

78

11

Phase I metabolites

ND

23

70

3

1.5.7.3  Conclusions

Phase II metabolites

ND

54

8

8

An understanding of species differences in the disposition of xenobiotics is of utmost importance, because the time course of dispositional events in an organism can be the crucial fac­ tor in the manifestation of toxicities. Thus, interpretation and,

a

Modified from Symchowicz et al. (1967). Expressed as percentage of dose; ND, not determined.

b

Table 1.20  Urinary, Biliary, and Fecal Excretion of Indomethacin and/or Its Metabolites in Dogs and Monkeys after IV Dosagea Compound

a

Urine

Bile

Dog

Monkey

Dog

Monkey

Dog

Monkey

Indomethacin

0.6b

10.5

3.8

33.6

68.7

4

Phase I metabolites

4.1

24.2

NI

NI

2.7

6

Phase II metabolites

3.3

17.9

52.1

8.1

3.1

NI

Total dose excreted

7.9

52.7

55.9

51.7

76.3

10

Modified from Hucker et al. (1966) and Yesair et al. (1970). Values represent percentage of dose excreted; NI, not identified or very small amounts.

b

Feces

Hayes’ Handbook of Pesticide Toxicology

88

more important, extrapolation of toxicity data from one spe­ cies to another is only possible if the kinetics of a xenobiotic are known. These discussions demonstrate that any and each step in the disposition of xenobiotics may be of major or minor importance for a particular compound in a particular species. However, there is sufficient knowledge available that an informed investigator can resolve the kinetic cause of spe­ cies differences for virtually any new chemical. More impor­ tant, only a broad view of the disposition of xenobiotics, as detailed in this overview, will enable a vigilant investigator to avoid the pitfalls of personal bias toward one or the other step as more or less important in the disposition of chemicals. Of utmost importance is knowledge of the half-life of pesticides for kinetic considerations in addition to the dose. Unfortunately, we do not have a compendium of critical data on pesticides as we do have it for drugs (Baselt, 2004).

1.5.8  Discussion of Factors Influencing Toxicity The fact that a number of factors influence dosageresponse relationships should not obscure the fact that these relationships are real and may be of importance. No special study seems to have been made of the inter­ relation of factors in regard to ratios of difference. If the highest ratios observed for a series of factors were com­ pletely multiplicative in their effect, the combined product would be very large. Although the possibility of such an occurrence cannot be excluded, none has been recognized. On the average, the ratios expressed as quotients differ lit­ tle from 1.0, and because some are less than 1.0, they tend to cancel out. Table 1.21 summarizes part of the information in the foregoing sections. It is clear from this summary and from

Table 1.21  Summary of Information on the Importance of Different Factors Influencing Toxicitya Factor

Duration Route Species

Total number of compounds 22 67

Ratio of difference Range

Increasing ratio indicatesb

Mean

0.5–20.0 c

0.2–21

—

2-year  90-day c

4.2 1

Oral  dermal

d

20

0.2–11.8

1

230

Other species  rat

1

1000

Human  cat

Oral route

69

1.20–7.14

2.42

LD 50  LD 1

Dermal route

42

1.37–14.93

3.00

LD 50  LD 1

Oral route

65

0.21–4.62

0.94

Male  female

Dermal route

37

0.11–2.93

0.81

Male  female

Pregnancy

19

0.74–14.55

1.90

Pregnant  nonpregnant

Age

18

0.6–10.0

2.9

Newborn  adult

16

0.7–6.2

—

Cold  warm

15

0.2–4.1

1.8

Other species  rat

Individuale

Sex

Newborn  adult f

290

0.02–750

2.78

Temperature

1

10,000

—

Cold  warm

Nutrition

8

1.0–1.8

1.49

1/3 dietary protein  normal No protein  normal

a

Expanded from Hayes (1967a), by permission of the Royal Society, London.  indicates greater toxicity of chemical or greater susceptibility of animal. Compounds with very low or variable toxicity are not included. d Approximate value. e Same sex. f Geometric mean. b c

Newborn  adult

Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

additional information in Tables 1.8 and 1.9 that species differences may be more important under practical condi­ tions than any factor except dosage and time in influenc­ ing toxicity of a particular compound. The largest ratio of difference found in connection with species was over 1000 whereas the largest ratio associated with any other fac­ tor likely to be of practical importance was only 21. It is true that very large ratios have been observed in connec­ tion with age and temperature, respectively, but their rarity must be emphasized. In summary, the maximal observed variation in effect associated with different factors is as follows: dosage and time-essentially infinite (health versus death) compounds107, temperature-104, age-103, species-103, other factors3  10 or less. The numerical comparison regarding species ignores important phenomena that occur in humans but are dif­ ficult or impossible to study in animals. One is forced to conclude that more emphasis should be placed on studies in humans. This is particularly true when one considers that not only dosage and time and route (which includes duration) but sometimes sex, age, temperature, duration of dosing, and other factors may be explored directly in vol­ unteers or workers. Compound (agent) and species (subject) determine the qualitative aspect of toxicity, whereas dose and time define the quantitative relationships of the interaction between subjects and agents. Dose- and time-dependences are mod­ ified by a multitude of factors discussed in this chapter, any one of which may under some circumstances impact on the c  t relationship. Relative potency is an intrinsic property of compounds often called a structure-activity relationship. Structure-activity relationships are mostly limited to closely related chemical structures because only compounds exerting the same effect (by the same mechanism of action) can be part of a structureactivity relationship (requirement for constancy of effect). Relative susceptibility is an intrinsic property of species for which the proper term in analogy to structure-activity would be species-reactivity. Species-reactivity relationships are often limited to closely related species because a given chemical (constancy of structure) will display a speciesreactivity relationship only as long as it is acting by the same mechanism. Coining these new terms instead of using the traditional notion of species differences was necessitated by a concept that has gone awry. The original purpose of studying species differences was to understand and know of the similarities between species because predictions can be based only on similarities but not on differences. These days species differences are being studied in a l’ars pour l’ars fashion which led many ignorant would-be toxicologists to claim that species-to-species predictions are impossible. As suggested by Hayes (1991) using limited human data obtained from volunteers or occupationally exposed workers in conjunction with detailed animal studies

89

conducted under ideal conditions (Sections 1.1.1–1.1.4) allows highly predictive safety-risk assessments when the principles of toxicology are applied according to the law(s) of toxicology.

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Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

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Great Britain Ministry of Agriculture, Fisheries and Food (1966). “Pesti­ cides Safety Precautions Scheme Agreed between Government Depart­ ment and Industry,” H. M. Stationery Office, London. Greenman, D. L., Bryant, P., Kodell, R. L., and Sheldon, W. (1982). Influence of cage shelf level on retinal atrophy in mice. Lab. Anim. Sci. 32, 353–355. Greenman, D. L., Kodell, R. L., and Sheldon, W. G. (1984). Association between cage shelf level and spontaneous and induced neoplasms in mice. J. Natl. Cancer Inst. 73, 107–113. Gunn, J. A., and Gurd, M. R. (1940). Action of some amines related to adrenaline: phenylallylamine, phenylbutenylamine, diphenylethyl­ amine. J. Physiol. (London) 98, 424–441. Guthrie, F. E., Monroe, R. J., and Abernathy, C. O. (1971). Response of the laboratory mouse to selection for resistance to insecticides. Toxicol. Appl. Pharmacol. 18, 92–101. Haber, F. (1924). On the history of gas warfare. In Five Lectures from the Years 1920-1923, pp. 76–92. Julius Springer, Berlin (in German). Hanson, H. B., and Hastings, A. B. (1933). The effect of smoking on the carbon monoxide content of blood. J. Am. Med. Assoc. 100, 1481. Harbison, R. D., and Becker, B. A. (1970). Effect of phenobarbital and SKF 525A pretreament on diphenylhydantoin teratogenicity in mice. J. Pharmacol. Exp. Ther. 175, 283–288. Hart, L. G., and Fouts, J. R. (1963). Effects of acute and chronic DDT administration on hepatic microsomal enzyme drug metabolism in the rat. Proc. Soc. Exp. Biol. Med. 114, 388–392. Hart, L. G., and Fouts, J. R. (1965). Further studies on the stimulation of hepatic microsomal drug metabolizing enzymes by DDT and its ana­ logs. Naunyn Schmiedebergs Arch. Exp. Pathol. Pharmacol. Hart, R., Fishbein, F., Hughes, D., and Smith, J. (1983). ED01 Workshop program: questions, issues and expectations. Fundam. Appl. Toxicol. 3, 127–198. Hartley, H. O., and Sielken, R. L. Jr. (1977a). Estimation of “safe doses” in carcinogenic experiments. Biometrics 33, 1–30. Hartley, H. O., and Sielken, R. L. Jr. (1977b). Estimation of “safe doses” in carcinogenic experiments. J. Environ. Pathol. and Toxicol. 1, 241–278. Hartung, R. (1987). Dose response relationships. In “Toxic Substances and Human Risk” (R. G. Tardiff and J. V. Rodricks, eds.), pp. 29–46. Plenum, New York. Haseman, J. K. (1985). Issues in carcinogenicity testing: dose selection. Fundam. Appl. Toxicol. 5, 66–78. Haseman, J. K. (1988). Lack of cage effect on liver tumor incidence in B6C3F1 mice. Fundam. Appl. Toxicol. 10, 179–187. Haseman, J. K., Crawford, D. D., Huff, J. E., Boorman, G. A., and McConell, F. F. (1984). Results from 86 two-year carcinogenicity studies conducted by the National Toxicology Program. J. Toxicol. Environ. Health 14, 621–639. Hatch, A., Balazs, T., Wiberg, G. S., and Grice, H. D. (1963). Long-term isolation stress in rats. Science 142, 407. Hayes, W. J. Jr. (1959). Pharmacology and toxicology of DDT. In DDT: The Insecticide Dichlorodiphenyltrichloroethane and Its Significance (P. Müller ed.), Vol. 2, pp. 9–247. Birkhäuser, Basel. Hayes, W. J. Jr. (1964). The toxicology of chemosterilants. Bull. W.H.O. 31, 721–736. Hayes, W. J. Jr. (1967a). Toxicity of pesticides to man: Risks from present levels. Proc. Roy. Soc. London, Ser. B. 167, 101–127. Hayes, W. J. Jr. (1967b). The 90-dose LD 50 and a chronicity factor as measures of toxicity. Toxicol. Appl. Pharmacol. 11, 327–335.

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Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

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Chapter | 1  Dose and Time Determining, and Other Factors Influencing, Toxicity

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parallel to a two-year chronic toxicity study. Fundam. Appl. Toxicol. 4, 663–670. Yeary, R., and Benish, R. A. (1965). A comparison of the acute toxicities of drugs in newborn and adult rats. Toxicol. Appl. Pharmacol. 7, 504.

Chapter 2

Pest Toxicology: The Primary Mechanisms of Pesticide Action John E. Casida University of California, Berkeley, California

2.1  Introduction Insects, plants, and fungi are our friends until they endanger our health and compete with us for food, at which time they become pests. Three millennia have past since Homer mentioned “pest-averting sulfur” and more than a century since dinitro-o-cresol became the first synthetic organic insecticide. With the introduction of the insecticide DDT, the herbicide 2,4-D, and the fungicide thiram in the 1930s and 1940s, the golden age of pesticide research (Casida and Quistad, 1998) and the chemical era of pest control were underway. The study of pesticides started largely as just “spray and count” for evaluating effectiveness. This era was quickly followed by curiosity and then a critical need for understanding what they do to people, crops, and the environment (pesticide toxicology and environmental toxicology) and how they work on the pest. The field of insect toxicology started by Hoskins in 1928 (Casida and Quistad, 2001) was soon expanded to include weeds and fungi. A new field was born to study how the pesticide works on the pest – that is, pest toxicology. To place pest toxicology in perspective as an aspect of pesticide science, policy, and management, the reader is referred to Tomlin (2006), Ware and Whitacre (2004), Stephenson and Solomon (2007), and Krieger (2009). Pesticides must be effective, selective, and safe. The benefits of pest management have to outweigh the economic, health, and environmental costs. Insecticides should be selectively toxic to pest insects compared with people and even relative to insect pollinators and beneficial arthropods. Herbicides designed to kill weeds must not harm closely related crops. Fungicides should control the grape disease fungus yet not interfere with the yeast fermentation to produce wine. The first generation of synthetic organic pesticides was generally used at 1–10 pounds per

Hayes’ Handbook of Pesticide Toxicology Copyright © 2009 American Chemical Society

acre. The effective doses for new compounds dropped 10- to 100-fold within the past half century. Pesticides are not only increasingly more potent but also of higher organismal specificity. The coupling of high potency with safety is achieved by utilizing unique differences at the target site level. Nature provides an amazing diversity of mechanisms for both pesticidal activity and selectivity. Species specificity is also sometimes dependent on pesticide metabolism (both activation and detoxification).

2.2  Primary targets Pesticides are intended to disrupt a primary target in the pest so it is no longer harmful. The pesticide per se or as its bioactivated form binds to or interacts with a specific enzyme, receptor, protein, or membrane, initiating a series of events that is deleterious or lethal to the pest. Insecticides and herbicides have between four and six primary targets that make up three-quarters of world sales (Figure 2.1). There are a few similar targets for the various pesticide types but they are usually very different. Most insecticides quickly disrupt neurotransmission to alter insect behavior or survival. Rapid action is usually required because insects cause economically important damage within a few hours or days. Insecticides can be practical with only a limited biological range like aphids or caterpillars. Herbicides generally inhibit plant-specific pathways, blocking amino acid or fatty acid biosynthesis or photosynthesis to “starve” the weed over several days. Fungicides act on many basic cellular functions important to hyphal tip growth (Figure 2.1). To be economically feasible they must control several diseases. Fungi are evolutionarily far more diverse than insects or weeds. They include not only the true fungi but also the Oomycetes having motile stages

103

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insecticide (mostly neurotransmission) other targets 1 2 3 4 5 6 7 8 9 10 11 12

5

chitin biosynthesis glutamate(chloride) receptor acetyl-CoA carboxylase ATP synthase ecdysone receptor uncoupler Bt toxin NADH dehydrogenase succinic dehydrogenase octopamine receptor unspecific unknown

6

8

7

9

10

11

12

4 3 2 acetylcholinesterase

1 GABAA receptor acetylcholine receptor

Voltage-gated sodium channel

herbicide (mostly plant specific pathways) other targets 1 2 3 4 5 6 7 8 9

tubulin photosystem I protoporphyrinogen IX oxidase 4-hydroxyphenyl pyruvate dehydrogenase phytoene desaturase glutamine synthase others unknown unspecific

2

4

3

5 6 7 8

9

1

EPSP synthase

acetyl-CoA carboxylase auxin receptor

acetolactate synthase fatty acid elongases

Photosystem II

fungicide (mostly basic cellular functions) other targets 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16

succinic dehydrogenase protein His kinase (osmo sensor) RNA polymerase scytalone dehydratase sterol ∆14 reductase uncoupler methionine biosynthesis protein kinase (osmo sensing) phospholipid biosynthesis protein biosynthesis (ribosomes) sterol 3-keto reductase

8

9 10

11

12 13

7

14 15

16

6 5

unspecific chemical reactives

4 3 2 1

tubulin

cytochrome c reductase

sterol c14α-demethylase

ATP synthase chitin biosynthesis dihydroorotate dehydrogenase inositol biosynthesis others, unknown

Figure 2.1  Insecticide, herbicide, and fungicide targets as percent of world sales shown as pie-shaped proportion of the total for 2003 (revised from Tietjen, 2003). Numbers designate minor targets in order of decreasing sales. “Others” include unspecific and unknown targets.

Chapter | 2  Pest Toxicology: The Primary Mechanisms of Pesticide Action

and controlled by oomyceticides. There are a broad variety of fungicide targets which vary in their importance for survival. As primitive microorganisms fungi are able to endure situations of energy depletion with the fungicide acting more as a fungistat while the disease is actually terminated by plant immune defense mechanisms.

2.3  Secondary targets The primary interaction usually occurs with the pesticide at picomolar or nanomolar levels and secondary interactions at higher concentrations. However, this is not always the case. There may be several targets of similar sensitivity but not equal importance. As an example, chlorpyrifos oxon inhibits not only acetylcholinesterase (AChE) but also several other serine hydrolases (Casida and Quistad, 2004), which although equally or more sensitive are not necessarily significant secondary targets. In addition, when studies are made in vitro at micromolar or millimolar levels, other secondary targets may become apparent although they are usually not toxicologically relevant in vivo compared to the primary effects. In yet another sense, lifetime feeding studies in rats, mice, and dogs at maximum tolerated doses, a critical part of the toxicology investigations for registration and establishing tolerance values, reveal biochemical and pathological changes not related to the primary target. These secondary targets in nonpest species play a major role in evaluating safety but are not considered further here where the focus is on the primary targets in the pest (i.e., pest toxicology).

2.4  Common target for structurally diverse pesticides The discovery of each new type of pesticide is followed by structural optimization for the highest possible potency involving sequential modification of each substituent to best fit the target site. This is often followed by a surge of activity by competitors to discover, develop, and patent an analog suitable to capture a portion of the potential market. The analog may have an advantage in cost-effectiveness, availability of intermediates, persistence, ease of metabolism, or safety. This is a predictable course of events. The surprise comes when a very different type of compound is found to work in the same way. Quite independent discoveries of the highly effective sulfonylurea and imidazolidinone herbicides were followed by the realization that they have the same primary target despite their very different structures. Some primary targets are highly specific in the ligands that bind while others have a broader scope for molecular recognition. Multiple classes of pesticides sometimes act at the same target as shown by competitive binding assays [e.g. respiratory inhibitors acting at the PSST site of Complex I (Schuler et al., 1999) and cyclodienes, fiproles, and picrotoxinin in blocking the -aminobutyric acid (GABA)-gated

105

chloride channel (Chen et al., 2006)]. The implications of the common target become particularly apparent relative to cross-resistance.

2.5  Resistance as a limiting factor Houseflies became resistant to DDT when continued use selected a strain with a less sensitive target site. The resistance to DDT extended to the pyrethrins and the synthetic pyrethroids. Cross-resistance was recognized as potentially a major problem. Some organophosphate (OP)-selected insects were resistant to both OPs and methylcarbamates (MCs). Weed resistance to atrazine conferred cross-resistance to some but not all herbicides acting at photosystem II (PSII). Fungi with target site resistance to triadimefon were not effectively controlled by some other C14demethylase inhibitors. The doses were elevated but this gave only a brief respite. All of the investment in developing a potent and safe pesticide can be lost without great care in managing the selection pressure during use. The severity of the problem was exacerbated by optimizing for target site potency, low doses, and specificity conferring safety, which ultimately favor the development of resistant strains based on a less sensitive primary target and more rapid metabolism of the exceedingly small amount of pesticide. Resistance conferred by detoxification has a completely different cross-resistance spectrum determined by metabolizable functional groups rather than target site insensitivity based on a common binding site. Pesticide management is a major aspect of pest control (i.e., reducing the selection pressure to slow resistance development). The rate of resistance development is dependent on the number of generations of pesticide selection per season which is normally 1 for plants compared to possibly 1–3 for insects and 12–25 for fungi. Resistance can be “disruptive” or “shifting” with very different consequences. Disruptive resistance with a factor of perhaps 1000 causes no fitness penalty for the pest. The pesticide does not work anymore and the resistant pest spreads and becomes established to eventually displace the wild type with continuing chemical use. By contrast, shifting resistance has a factor of maybe 2–10 and is connected with a fitness penalty. A higher dose of pesticide is required but at the end of pesticide use the population shifts back to the wild type. When resistance appears to one target site or mode of action, the pesticide is replaced by another one with a different mode of action or resistance group. The importance of pesticide management led to the formation of the Insecticide Resistance Action Committee (IRAC) (2008), the Herbicide Resistance Action Committee (HRAC) (2005), and the Fungicide Resistance Action Committee (FRAC) (2007). The very knowledgeable experts on these committees are mostly from industries involved in pesticide research and development. Their compilations defining resistance groups

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are actually the outlines and listings of primary target sites in pest toxicology. The discussion below of comparative biochemistry or molecular toxicology is based largely on the IRAC, HRAC, and FRAC reports considered by primary target sites rather than type of pesticides.

Target site resistance can be a major limiting factor for insecticide action at a common nerve target – that is, the OPs and MCs at AChE, the pyrethroids and DDT at the sodium channel, and the cyclodienes and phenylpyrazoles at the GABA-gated chloride channel. Some relief is provided by indoxacarb and the avermectins.

2.6  Nerve (Table 2.1, Figure 2.2) Most insecticides by number, amount, and market value act on the nervous system at the synapse or the axon. The cholinergic system is the principal insecticide target with OP and MC compounds inhibiting AChE to prolong the excitatory action of acetylcholine (ACh). The nicotinic ACh receptor (nAChR) is the target for the neonicotinoids as competitive agonists for ACh, spinosad as an allosteric modulator, and cartap as a noncompetitive antagonist. The axonal voltage-gated sodium channel is the target of DDT, pyrethrins, and pyrethroids acting as modulators and indoxacarb as a blocker. Synaptic neurotransmission at the GABA-gated chloride channel is the target for the noncompetitive antagonists and blockers endosulfan and fipronil and the GABA/glutamate-gated chloride channel is stimulated by the avermectins. The G protein, coupled octopamine receptor is the target for the agonist amitraz.

Figure 2.2  Most insecticides act on the nervous system at several synaptic and axonal sites.

Table 2.1  Nerve Targets for Insecticides System

Compounds

Chemical type Example

Number

Organophosphate Methylcarbamate

Chlorpyrifos Carbaryl

64 26

   a. Competitive agonist

Neonicotinoid Botanical alkaloid

Imidacloprid Nicotine

7 1

   b. Allosteric agonist

Spinosyn

Spinosad

2

   c. Antagonist

Nereistoxin analog

Cartap

4

  1. Modulator

Pyrethroid DDT analog

Pyrethrins DDT

31 2

  2. Voltage-gated sodium channel blocker

Oxadiazine Semicarbazone

Indoxacarb Metaflumizone

1 1

  1. GABA-gated chloride channel antagonist

Cyclodiene Phenylpyrazole

Endosulfan Fipronil

2 2

  2. Glutamate-gated chloride channel activator

Avermectin

Abamectin

3

D. Octopamine receptor agonist

Formamidine

Amitraz

1

A. Cholinergic   1. Acetylcholinesterase   2. Nicotinic acetylcholine receptor

B. Sodium channel

C. Chloride channel

Chapter | 2  Pest Toxicology: The Primary Mechanisms of Pesticide Action

2.7  Photosynthesis and pigment synthesis (Table 2.2, Figure 2.3) Green plant pigments absorb light and with the coupled systems of chloroplasts convert light energy to the chemical energy of adenosine triphosphate (ATP). Herbicides disrupting these processes unique to plants are usually of low toxi­ city to mammals which lack analogous targets. PSII was an early target for herbicides and is still highly important, being acted upon by 50 commercial compounds. More than one target is involved since resistance to one PSII inhibitor

107

does not confer cross-resistance to all others denoted here as the triazine, urea, and nitrile “sites.” The photosystem I (PSI) electron pathway is diverted by bipyridilium herbicides with paraquat as the principal example. Protoporphyrinogen IX oxidase is the target for 26 herbicides of many chemical types. Carotenoids protect chlorophylls from overactivation and destruction by light. These yellow/ orange pigments must be present to protect the green pigments. Inhibition of four herbicide targets leads to bleaching and weed death. Phytoene desaturase is highly sensitive to seven herbicides mostly with m-trifluoromethylphenyl

Table 2.2  Photosynthesis and Pigment Synthesis Targets for Herbicides System

Compounds

Chemical type Example

Number

A. Photosynthesis   1. PSII    a. Triazine site

Triazine Triazinone Uracil Phenylcarbamate Triazolinone Pyridazinone

Atrazine Metribuzin Bromacil Desmedipham Amicarbazone Pyrazon

14 3 3 2 1 1

   b. Urea site

Urea

Diuron

18

   c. Nitrile site

Amide Nitrile Phenylpyridazine Benzothiadiazinone

Propanil Bromoxynil Pyridate Bentazon

2 3 2 1

  2. PSI electron diversion

Bipyridylium

Paraquat

2

  3. Protoporphyrinogen IX oxidase

Diphenylether N-phenylphthalimide Triazolinone Phenylpyrazole Thiadiazole Oxadiazole Pyrimidindione Oxazolidinedione Other

Acifluorfen-Na Flumiclorac-pentyl Azafenidin Fluazolate Fluthiacet-methyl Oxadiazon Benzfendizone Pentoxazone Pyrafluazol

8 3 3 2 2 2 2 1 3

  1. Phytoene desaturase

Pyridinecarboxamide Pyridazinone Other

Diflufenican Norflurazon Fluridone

2 1 4

  2. Lycopene cyclase

Triazole

Amitrole

1

  3. 4-Hydroxyphenyl pyruvate dehydrogenase

Pyrazole Triketone Isoxazole Other

Benzofenap Mesotrione Isoxachlortole Benzobicyclon

3 2 2 1

  4. Unknown

Isoxazolidinone Urea Diphenylether

Clomazone Fluometuron Aclonifen

1 1 1

B. Pigment synthesis (bleaching)

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substituents. Lycopene cyclase is inhibited by amitrole. 4-Hydroxyphenyl pyruvate dehydrogenase inhibition by eight herbicides leads to bleaching by an entirely different sequence of reactions. There are also three other bleachers of different chemical types and unknown target.

2.8  Biosynthesis 2.8.1  Herbicides (Table 2.3, Figure 2.4) Figure 2.3  Photosystem II is inhibited by a great variety of herbicides and the photosystem I electron pathway is diverted by bipyridilium compounds.

Plants synthesize their own amino acids whereas animals do not and require in their diet the essential amino acids

Table 2.3  Biosynthesis Targets for Herbicides Target

Compounds

Chemical type Example

Number

A. Amino acid   1. EPSP synthase

Glycine derivative

Glyphosate

2

  2. AHAS or ALS

Sulfonylurea Imidazolinone Triazolopyrimidine Pyrimidinyloxybenzoate Sulfonylaminocarbonyl-triazolinone

Chlorsulfuron Imazapyr Flumetsulam Bispyribac-Na Flucarbazone-Na

31 6 6 5 2

  3. Glutamine synthase

Phosphinic acid

Bialaphos

2

  1. Microtubule assembly

Dinitroaniline Phosphoroamidate pyridine Benzamide Benzoic acid

Trifluralin Butamiphos Dithiopyr Propyzamide Chlorthal-dimethyl

7 2 2 2 1

  2. Mitosis/microtubule organization

Carbamate

Chlorpropham

3

  3. DHP synthase

Carbamate

Asulam

1

  1. Acetyl-CoA carboxylase (ACCase)

Cyclohexanedione Aryloxyphenoxy-propionate Phenylpyrazoline

Sethoxydim Diclofop-methyl Pinoxaden

8 8 1

  2. Not ACCase

Thiocarbamate Chlorocarbonic acid Benzofuran Phosphorodithioate

Molinate Dalapon Benfuresate Bensulide

14 3 2 1

  3. Very long chain fatty acid synthesis

Chloroacetamide Acetamide Oxyacetamide Tetrazolinone Other

Acetochlor Diphenamid Flufenacet Fentrazamide Anilofos

12 3 2 1 3

D. Cell wall (cellulose)

Nitrile Benzamide Triazolocarboxamide

Dichlobenil Isoxaben Flupoxam

2 1 1

B. Microtubule and cell division

C. Fatty acid synthesis

Chapter | 2  Pest Toxicology: The Primary Mechanisms of Pesticide Action

that they cannot make. Amino acid biosynthesis is therefore a preferred target for herbicides since there are no corresponding systems in mammals. Three major targets are involved: enolpyruvylshikimate 3-phosphate synthase (EPSP) for glyphosate and one other herbicide; acetohydroxy acid synthase (AHAS), also known as acetolactate synthase (ALS), for 31 sulfonylureas, six imidazolinones, and 13 chemicals of other types; and glutamine synthase for glufosinate and one other compound. Overexpressed EPSP synthase and glutamine synthase as low sensitivity targets are normally coupled with overexpressed herbicide detoxification systems for enhancing crop tolerance. The microtubule system and cell division have three herbicide targets. The largest number of compounds (17) of varied chemical type including trifluralin alter the microtubule assembly process. Three carbamates inhibit mitosis/microtubule organization and another dihydropteroate (DHP) synthase.

Figure 2.4  Plant amino acid biosynthesis is inhibited at the most important EPSP synthase and AHAS sites and two other enzymes by many herbicides.

109

Fatty acid synthesis is a favored target as evident from the 58 herbicides acting in this way, some on acetyl-CoA carboxylase (ACCase) and others at diverse sites altering very long chain fatty acid synthesis. Cell wall biosynthesis is inhibited by four herbicides including dichlobenil.

2.8.2  Fungicides and Insecticides (Table 2.4, Figure 2.5) The fungal sterol is ergosterol versus the mammalian cholesterol with critical differences in their biosynthetic pathways that allow selective inhibition. Four sterol targets are involved in fungicide action. The most important is the C14-demethylase inhibited by 26 triazoles including triadimefon, five imidazoles, and four other fungicides. This site is highly sensitive and very prone to target site insensitivity, but fortunately the selection is counteracted somewhat by the fitness penalty resulting in only shifting resistance. The 14-reductase and 8→7-isomerase are inhibited by four morpholines and some structurally unrelated compounds and two other steps by diverse chemicals. Four targets of fungicides in nucleic acid biosynthesis are RNA polymerase I (six fungicides, including benalaxyl), adenosine-deaminase (three fungicides), DNA/RNA synthesis (two fungicides), and DNA topoisomerase type II (gyrase) (one compound). Four antibiotic fungicides such as streptomycin block protein synthesis. Antitubulin fungicides mostly affect -tubulin assembly in mitosis (eight compounds of diverse types with benomyl as an example) while others block cell division or delocalize spectrin-like proteins (two compounds). Phospholipid biosynthesis is blocked by four fungicides inhibiting the methyl transferase while isoprothiolane and five others disrupt cell wall

Table 2.4  Biosynthesis Targets for Fungicides (f) and Insecticides (i) Target

Chemical type

Compounds Example

Number

A. Sterol (f )   1. C14-demethylase

Triazole Imidazole Pyrimidine Piperazine Pyridine

Triadimefon Imazalil Fenarimol Triforine Pyrifenox

26 5 2 1 1

  2. 14-reductase and 8→7-isomerase

Morpholine Piperidine Spiroketal-amine

Adimorph Fenpropidin Spiroxamine

4 2 1

  3. 3-keto reductase, C4-demethylation

Hydroxyanilide

Fenhexamid

1

  4. Squalene-epoxidase

Allylamine Thiocarbamate

Naftitine Pyributicarb

2 1 (Continued)

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Table 2.4  (Continued) Target

Chemical type

Compounds Example

Number

B. Nucleic acid (f )   1. RNA polymerase I

Acylalanine Oxazolidinone Butyrolactone

Benalaxyl Oxadixyl Ofurace

4 1 1

  2. Adenosine-deaminase

Hydroxy-(2-amino) pyrimidine

Buprimate

3

  3. DNA/RNA synthesis (proposed)

Isoxazole Isothiazolone

Hymexazole Octhilinone

1 1

  4. DNA topoisomerase type II (gyrase)

Carboxylic acid

Oxolinic acid

1

C. Protein (f )

Antibiotic

Blasticidin-S Kasugamycin Streptomycin Oxytetracycline

1 1 1 1

  1. -tubulin assembly in mitosis

Benzimidazole Thiophanate

Benomyl Thiophanate-methyl

4 2

  2. -Tubulin assembly in mitosis

N-phenylcarbamate

Diethofencarb

1

  3. -Tubulin assembly in mitosis

Toluamide

Zoxamide

1

  4. Cell division (proposed)

Phenylurea

Pencycuron

1

  5. Delocalization of spectrin-like proteins

Pyridinylmethylbenzamide

Fluopicolide

1

   a. P  hospholipid biosynthesis methyl transferase (f )

Phosphothiolate Dithiolanes

Iprobenfos Isoprothiolane

3 1

   b. Phospholipid biosynthesis and cell wall deposition (f )

Carboxylic acid amide

Dimethorph

6

   c. Lipid biosynthesis (i)

Tetronic acid

Spiromesifen

2

  2. Lipid peroxidation (f )

Amino/nitrobenzene Benzene Thiophosphate Thiadiazole

Dicloran Biphenyl Tolclofosmethyl Etridiazole

3 2 1 1

  3. Cell membrane permeability, fatty acid (f )

Amides and thioamides

Prothiocarb

3

  1. Trehalase and inositol

Glucopyranosyl antibiotic

Validamycin

1

  2. Chitin synthase

Peptidyl pyrimidine nucleoside

Polyoxin

1

G. Methionine biosynthesis (f )

Aniline-pyrimidines

Cyprodinil

3

  1. Reductase

Various

Tricyclazole

3

  2. Dehydratase

Carboxamides

Carpropamid

3

D. Mitosis and cell division (f )

E. Lipid   1. Lipid and phospholipid biosynthesis

F. Glucans (f)

H. Cell wall (melanin) (f )

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deposition. Lipid peroxidation is initiated by seven fungicides and cell membrane permeability is disrupted by three others. Biosynthesis of glucans is inhibited by two fungicides (including polyoxin for chitin synthase), of methionine by cyprodinil and two others, and of melanin by six fungicides. Insecticidal activity is achieved with tetronic acids, such as spiromesifen, inhibiting lipid biosynthesis.

2.9  Respiration (Table 2.5, Figure 2.6)

Figure 2.5  Fungal ergosterol biosynthesis is inhibited at the most important CYP450 C14-demethylase and three other sites by a great variety of fungicides.

The synthesis of ATP, the energy currency of the cell, is a complex process carried out by the mitochondrial respiratory electron transport chain involving a series of five membrane-bound complexes (I–V). Pesticides disrupt many sites by binding and inhibition (I–IV) or acting as

Table 2.5  Respiration Targets for Insecticides (i), Herbicides (h), and Fungicides (f) Target

Chemical type

Compounds Example

Number

A. Electron transport   1. Complex I NADH oxidoreductase    a. Coupling site I (i)

Various

Rotenone

7

   b. Other (f)

Pyrimidinamine

Diflumetorim

1

Carboxamide Various Acrylonitrile

Carboxin Penthiopyrad Cyenopyrafen

1 9 1

   a. Coupling site 2 (i)

Various

Acequinocyl

3

   b. Ubiquinol oxidase at Qo site (f��� ��)

Various Methoxyacrylate

Kresoximmethyl Azoxystrobin

10 3

   c. Cytochrome b Qo site (i)

Carbazate

Bifenazate

1

   d. Cytochrome bc1 at Qi site (f��� ��)

Cyanoimidazole Sulfamoyltriazole

Cyazofamid Amisulbrom

1 1

  4. Complex IV (i)

Fumigant

HCN, PH3

3

  1. Uncouplers via disruption of proton gradient (i, h, f��� ��)

Dinitrophenol Liposoluble insecticide Dinitroaniline Pyrimidinone-hydrazone

Binapacryl Chlorfenapyr Fluazinam Ferimzone

7 1 1 1

  2. ATP synthase (i)

Carbodiimide progenitor

Diafenthiuron

1

  3. ATP production (i, f��� ��)

Triorganotin Thiophenecarboxamide

Cyhexatin Silthiofam

6 1

C. Aconitase (i)

Haloaliphatic acid

Fluoroacetate

1

D. Others (i)

Various

Propargite

2

  2. Complex II    a. Succinic dehydrogenase (i,f )

  3. Complex III

B. Oxidative phosphorylation

112

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Figure  2.6  The mitochondrial respiratory electron transport chain has pesticide inhibition sites in each of the five membrane-bound complexes (modified from Schuler and Casida, 2001).

uncouplers to prevent oxidative phosphorylation and formation of the proton gradient. Rotenone and a series of miticides inhibit by binding to the PSST site in Complex I. Carboxin and nine other fungicides and a recently reported metabolite of the acaricide cyenopyrafen inhibit succinic dehydrogenase in Complex II, and the strobilurins block the quinol oxidation center of Complex III. The insecticide acequinocyl inhibits Complex III coupling site 2, bifenazate at cytochrome b Qo site, and two fungicides at the cytochrome bc1 Qi site. Cyanide and phosphine block Complex IV and carbodiimides and triorganotins the ATP synthase of Complex V. The insecticide chlorfenapyr is one of 10 pesticidal uncouplers of oxidative phosphorylation. Fluoroacetate is converted to fluorocitric acid which inhibits aconitase of the tricarboxylic acid cycle. The steps in electron transport are sufficiently conserved between insects and mammals that it is difficult to achieve large degrees of selectivity for inhibitors. In animals inhibition of the respiratory chain leads to radical accumulation and induces apoptosis, which is especially damaging in neuronal cells. Inhibition of the respiratory chain in plants can be irrelevant as long as photosynthesis supplies NADH and ATP and in fungi it leads to a type of starvation.

in time and amount of juvenile hormone to stay young, and growth and differentiation hormone or ecdysone to develop, molt, and become an adult. Juvenile hormone mimics and analogs such as methoprene are very effective and selective but provide slow control. Molting disruptors include ecdysone receptor agonists, e.g. diacylhydrazines such as tebufenozide that act at the ecdysone binding site. Cryomazine for Diptera is also a molting disruptor as is the natural azadirachtin by undefined mechanisms. Chitin biosynthesis inhibitors for Lepidoptera (benzoylphenyl ureas) and Homoptera (buprofezin) act in very different ways. The benzoylphenyl ureas block chitin synthesis in vivo but not chitin synthase in vitro, so the mechanism is unsolved. In comparison, cellulose biosynthesis-inhibiting herbicides (dichlobenil) and cell wall biosynthesis-inhibiting oomyceticides (iprovalicarb and mandipropamid) also do not directly inhibit the sugar polymerase. Plant disease development is regulated by auxins, ethylene, cytokinins, and gibberellins, among others. Several types of carboxylic acids such as 2,4-D serve as synthetic auxins and other compounds including naptalam as auxin transport inhibitors. Finally, fungal growth can be altered by host defense inducers such as acibenzolar S-methyl and probenazole.

Figure 2.7)

2.10  Growth regulators (Table 2.6,

2.11  Unknown, nonspecific   and other targets (Table 2.7)

Every organism follows a programmed course of growth and development carefully synchronized for species propagation and environmental integration. Compounds that disrupt these delicate hormone-guided processes serve as insect growth regulators (IGRs), plant growth regulators (PGRs), and fungal growth regulators or host plant defense inducers. Insect development is controlled by a balance

The endotoxins of the bacterium Bacillus thuringiensis (Bt) disrupt insect midgut membranes with effectiveness dependent on the strain of Bt and the particular pest. “Bt crops” with expressed recombinant endotoxin such as Cry1Ab play a critical role in control of lepidopterous larvae. The newly introduced diamide ryanodine receptor modulators offer great promise based on potency for lepidopterous

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113

Table 2.6  Growth Regulator Targets Organism

Chemical type

Compounds Example

Number

Juvenile hormone analog Phenoxyphenoxy Ether

Methoprene Fenoxycarb

3 2

   a. Ecdysone agonist

Diacylhydrazine

Tebufenozide

4

   b. Dipteran

Triazine

Cyromazine

1

   c. Botanical

Neem constituent

Azadirachtin

1

   a. Lepidopteran

Benzoylurea

Diflubenzuron

11

   b. Homopteran

Buprofezin

Buprofezin

1

  1. Synthetic auxin

Phenoxy carboxylic acid Pyridine carboxylic acid Benzoic acid Quinoline carboxylic acid Other

2,4-D Picloram Chloramben Quinclorac Benazolin-ethyl

8 4 3 2 1

  2. Auxin transport

Semicarbazone Pthalamate

Diflufenzopyr-Na Naptalam

1 1

  3. Ethylene generator

Chloroethylphosphonic acid

Ethephon

1

  4. Cytokinin

Adenine derivatives

Benzyladenine

2

  5. Gibberellin

Diterpenoid acid

Gibberellic acid

2

  1. Salicylic acid pathway

Benzothiadiazole BTH

Acibenzolar S-methyl

1

  2. Other

Thiadiazole-carboxamide Benzisothiazole Natural

Tiadinil Probenazole Laminarin

2 1 1

A. Insect and IGRs   1. Juvenile hormone mimic   2. Molting disruptor

  3. Chitin biosynthesis

B. Plant and PGRs

C. Fungi-host plant defense inducers

larvae and safety. One or both of the selective feeding blockers (pymetrozine and flonicamid) may disrupt nerve processes. Compounds that inhibit mite growth often differ in mode of action from those for insects but the mechanisms remain unknown. Synergists act both as CYP450 monooxygenase inhibitors such as piperonyl butoxide and esterase inhibitors with tribufos as an example. Methyl bromide continues to be used as a major fumigant. There are six other unknown targets for insecticides. The herbicide endothal inhibits protein phosphatase 2A, while 19 other herbicides act on unknown targets. Seven fungicides disrupt signal transduction at two different sites. The multisite fungicides have been extremely important for many decades with dithiocarbamates, copper, and sulfur as major examples. There are also 16 fungicides of unknown mode of action.

2.12  Overview (Table 2.8) A compilation of pesticides by targets (including different binding sites in the same target) and numbers based on Tables 2.1–2.7 is presented as an overview in Table 2.8. It is no surprise that only the insecticides act on nerve processes and only the herbicides on photosynthesis and pigment synthesis light processes. Biosynthesis inhibitors include major herbicides and fungicides in number and variety of structures. Respiration inhibitors are largely insecticides and fungicides and the growth regulators are particularly important for insects and plants. There are a large number of “other” targets, often with only a single example. In this form of compilation, the numbers of targets are similar for insecticides, herbicides, and fungicides, whereas the number of compounds is less for fungicides.

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Figure 2.7  Insect growth regulators disrupt the molting process by acting as juvenile hormone mimics (e.g. methoprene), ecdysone receptor agonists (e.g. tebufenozide), or chitin biosynthesis inhibitors (e.g. diflubenzuron).

Table 2.7  Unknown, Nonspecific and Other Targets Chemical type

Compounds Example

Number

A. Insecticide   1. M  icrobial disruptor of insect midgut membrane

Bacillus thuringiensis (Bt) Crop proteins

Cry1Ab

8

  2. Ryanodine receptor modulators

Diamide

Flubendiamide

2

  3. Selective feeding blocker

Various

Pymetrozine

2

  4. Mite growth inhibitor

Various

Clofentezine

3

   a. Cyp450 monooxygenase

Methylenedioxyphenyl

Piperonyl butoxide

1

   b. Esterase

OP

Tribufos

1

  6. Fumigant

Alkyl halide

Methyl bromide

3

  7. Other

Benzoximate

Benzoximate

6

  1. Protein phosphatase 2a

Dicarboxylic acid

Endothal

1

  2. Unknown

Arylaminopropionic acid Organoarsenical Pyrazolium Other

Flamprop-Mmethyl DMSA Difenzoquat Cinmethylin

2 2 1 14

  5. Synergists

B. Herbicide

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115

Table 2.7  (Continued) Organism

Chemical type

Compounds Example

Number

C. Fungicide   1. Signal transduction

Quinoline

Quinoxyfen

1

   a. G proteins in early cell signaling

Dicarboximide

Vinclozolin

4

   b. Map/histidine kinase

Phenylpyrrole

Fenpiclonil

2

  2. Multisite

Dithiocarbamate Phthalimide Chloronitrile Various organic Various

Maneb Captan Chlorothalonil Dodine Inorganic Cu, S

8 3 1 7 2

  3. Unknown

Various Diverse

Fosetyl-Al Mineral oils

12 4

Table 2.8  Number of Pesticide Targets Including Different Binding Sites and Compounds Acting at Those Targets Number of targetsa

Type of target

Number of compoundsb

Insect

Herb

Fung

Total

Insect

Herb

Fung

Total

Nerve

9

0

0

9

147

0

0

160

Photosynthesis and pigment synthesis

0

9

0

9

0

97

0

97

Biosynthesis

1

10

23

34

2

134

103

239

Respiration

10

1

6

17

36

10

43

89

Growth regulators

6

5

2

13

23

25

5

53

Others

7

2

4

13

26

20

44

88

Total

33

27

35

95

234

286

195

715

a

Includes unknown targets tabulated as 1 in each category.

b

The same compound may appear in two or three categories if it has multiple uses or actions such as pesticides affecting oxidative phosphorylation.

A totally different picture of pesticides is obtained on evaluating by percent of world market value or amount used. Glyphosate with glyphosate-resistant crops is the most important by world market value and sulfur by amount used in California. Several aspects of pest management are not considered. These include insect pheromones, nematicides, and rodenticides. This overview is dependent primarily on the IRAC, HRAC, and FRAC sources used,

and secondarily on the approach in arranging and interpreting the available information. The author proposed about 20 years ago that there may be a finite number of practical targets for pesticide action (Casida, 1990). The frequency at which the new compounds acted at the same old targets raised concerns that we cannot indefinitely rely on discovering new targets. However, screening of natural products and synthetic

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compounds has continued to come up with a diversity of mechanisms and effective structures. There has also been a concerted search for new types of pest control agents that ultimately led to new targets. Clearly there are many more yet to discover. However, despite many advances in recent years, the number of targets with more than 5% market share in 2003 was only four for insecticides, six for herbicides, and four for fungicides (Figure 2.1) (see also Tietjen et al., 2005).

Conclusion Pests are currently controlled with about 715 pesticides acting by perhaps 95 different mechanisms. Can we stop research now and just continue to use the current compounds effectively and safely? Do we really need more pesticides? Advances in chemical pest control are dependent on a thorough understanding of both pest and pesticide toxicology. We must be ready when pests inevitably develop resistance. With current pesticides, pests are generally much more sensitive than people and crops. However, we can still achieve higher target site potency and lower use levels, reduced environmental impact, and improved safety. Pesticide legislation policies for risks and hazards continue a trend toward ever-greater restrictions. We must constantly maintain a pool of safe pesticides acting on many targets to meet demands for increased food production and improved human health. Continued success in pest management depends on developing and using an expanding knowledge base in comparative biochemistry and molecular toxicology considering pests, people, and crops. The time has arrived when advances in pesticide toxicology are dependent on or coupled with discoveries in pest toxicology.

Postscript The following paragraph is a comment on “Why I do Science” (Casida, 2008): I find that crossword and jigsaw puzzles are not so much fun because you know there is a solution. But try a mechanism study on a chemical that works on a totally unknown target—now you can really have fun. It may be easy, or it may take 50 years before the background science is available. Alternatively, the new chemical and novel mode of action serves as a probe to dissect a new area of science. Pesticides—chemicals that affect the growth or survival of a pest—are particularly intriguing. Hundreds of thousands of compounds are sifted every year in the search for very unusual effects, which are great fun to sort out. Each new discovery, advance, or stage of understanding carries with it the thrill of the moment, but they are really for all time as they become part of our knowledge base and toolkit.

The approach of our laboratory is to take a new compound and optimize its potency—meaning that you can use less of it because it is designed to go to just the right place to do the assigned job. When other strategies fail, we use tritium to make the compound highly radioactive and then use this radioligand to quantitate, assay, purify, isolate, and ultimately identify the target. Genomics, proteomics, and all the other “omics” really help us solve problems as never before. Once you identify the mechanism, you can manipulate a life process. Can you create a useful new pesticide, a new cancer drug, or a new way to measure and modulate a receptor in the brain? Then the challenge is to find a way to use the chemical without side effects while fitting the economic reality of the marketplace. We leave that to the entrepreneurs.

Acknowledgments The author gives special thanks to Alex Gulevich who assisted with devotion and distinction in compiling and presenting the information; to Motohiro Tomizawa, Ralf Nauen, Dale Shaner, and Klaus Tietjen who provided helpful review comments; and to the University of California at Berkeley William Muriece Hoskins Chair in Chemical and Molecular Entomology for continuing support.

References Casida, J. E. (1990). Pesticide mode of action: evidence for and implications of a finite number of biochemical targets. In “Pesticides and Alternatives: Innovative Chemical and Biological Approaches to Pest Control” (J. E. Casida ed.), pp. 11–12. Elsevier, Amsterdam. Casida, J. E. (2008). Why I do science. In “Breakthroughs.” College of Natural Resources, University of California, Berkeley, Summer issue, p.6. Casida, J. E., and Quistad, G. B. (1998). Golden age of insecticide research: past, present, or future? Annu. Rev. Entomol. 43, 1–16. Casida, J. E., and Quistad, G. B. (2001). Insect toxicology in the beginning: William Muriece Hoskins. Pest Manag. Sci. 57, 875–876. Casida, J. E., and Quistad, G. B. (2004). Organophosphate toxicology: safety aspects of non-acetylcholinesterase secondary targets. Chem. Res. Toxicol. 17, 983–998. Chen, L., Durkin, K. A., and Casida, J. E. (2006). Structural model for -aminobutyric acid receptor noncompetitive antagonist binding: widely-diverse structures fit the same site. Proc. Natl. Acad. Sci. USA 103, 5185–5190. Fungicide Resistance Action Committee (2007). FRAC code list: fungicides sorted by mode of action. http://www.frac.info/frac/publication/ anhang/FRAC_Code_List_2007_web.pdf Herbicide Resistance Action Committee (2005). Classification of herbicides according to mode of action. http://www.hracglobal. com/Publications/ClassificationofHerbicideModeofAction Insecticide Resistance Action Committee (2008). IRAC mode of action classification version 6.1. http://www.irac-online.org Krieger, R. I. (ed.) (2009). “Handbook of Pesticide Toxicology,” 3rd ed. Academic Press, San Diego. Schuler, F., and Casida, J. E. (2001). The insecticide target in the PSST subunit of complex I. Pest Manag. Sci. 57, 932–940.

Chapter | 2  Pest Toxicology: The Primary Mechanisms of Pesticide Action

Schuler, F., Yano, T., Di Bernardo, S., Yagi, T., Yankovskaya, V., Singer, T. P., and Casida, J. E. (1999). NADH-quinone oxidoreductase: PSST subunit couples electron transfer from iron-sulfur cluster N2 to quinone. Proc. Natl. Acad. Sci. USA 96, 4149–4153. Stephenson, G. R., and Solomon, K. R. (2007). “Pesticides and the Environment.” Canadian Network of Toxicology Centres Press, Guelph, Ontario, Canada. Tietjen, K. (2003). Mode of action research for novel crop protection products. Oral presentation.

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Tietjen, K., Drewes, M., and Stenzel, K. (2005). High throughput screening in agrochemical research. Comb. Chem. High Throughput Screen. 8, 589–594. Tomlin, C. D. S. (ed.) (2006). “The Pesticide Manual,” 14th ed. British Crop Protection Council, Farnham, UK. Ware, G. W., and Whitacre, D. M. (2004). “The Pesticide Book.” MeisterPro Information Resources, Willoughby, OH.

Chapter 3

Pest Control Agents from Natural Products István Ujváry iKem BT, H-1033 Budapest, Hungary Besides, hellebore is rank poison to us, but given to goats and quails makes them fat. Lucretius: De Rerum Natura, Book 4, lines 640–641 Translated by W. H. D. Rouse

3.1  Introduction Natural products have been used to control animal pests, plant diseases, and weeds since ancient times. Plants have been the most important sources of natural pesticides for centuries and preparations standardized for the active ingredient(s) by modern analytical methods became available in the past decades, making possible the manufacture of reliable products. More recently the immense potential of bacteria and other microorganisms for the production of biologically active substances was realized and many new pest control agents commercialized since the middle of the 20th century are of microbial origin. Living organisms generally produce a mixture of structurally related compounds, of which often one or two are found in abundance. Small structural changes, brought about by the intricate web of metabolic processes of the producing organism, are usually reflected in variations in the pharmacological, including toxicological and pharmacokinetic characteristics of the compounds. Moreover, the combined effect of the constituents of a crude preparation often complicates the evaluation of the mixture. Minor components can have unique, either favorable or unfavorable, biological properties that are unveiled only after separation of the ingredients. Although crude or partly purified preparations still continue to be used in practice, reliability and safety dictates the use of pure or at least enriched and, if necessary, stabilized products. By now, the main active ingredients of all currently used natural pest control agents have been isolated and characterized (reviewed by Copping and Duke, 2007; see also Copping, 2004; Ujváry, 2003). The true ecological function of these bioactive substances in the producing organism has rarely been clarified, although a defensive role is generally assumed. The complexities of the underlying biochemistry and genetics are better understood and this facilitates the development of strains that produce a desired bioactive natural product in an economically feasible way. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

Clarification of the mode of action of essentially all of these compounds at the receptor level was made possible by advances in molecular biology during the past three decades. This information is also very helpful in the prevention and treatment of poisoning. The biological activity spectrum of natural pesticides is extremely variable, and the selectivity is often achieved by targeted application, often relying on the behavior of the target pest. Among the substances to be discussed, one finds the botanical insecticide pyrethrum, with an unparalleled record of safe use; the extremely poisonous alkaloid strychnine, the use of which is severely restricted; and the tobacco alkaloid nicotine, which is currently of limited use as an insecticide but is generally accepted by hundreds of millions as a recreational, though addictive, drug. Natural compounds are not necessarily safer than synthetic ones. As Coats (1994) has recently summarized: 1. The biological property of a chemical is a function of its structure rather than its origin. 2. The biological activity exerted by a given dose of the compound under given circumstances, especially as related to safety, depends on the way in which the chemical is used. Relying on the results of carcinogenicity studies with hundreds of natural and synthetic chemicals, Ames et al. (1990) (see also Gold et al., 2001) have also refuted the general (and popular) assertion that “Natural is safe.” However, natural substances used for pest and disease control are generally nonpersistent under field conditions. Most of these often-complex molecules are readily transformed abiotically by light and/or oxygen into less toxic products. There is also extensive biotransformation occurring in the soil, water, and plants to which these compounds are applied so no residues are expected. For most traditional natural pesticides the acute toxicity data that have accumulated over the past 50–100 years provide useful information on the risks associated with regular exposure during application and from residues. There 119

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is, of course, a substantially higher risk associated with the production, transportation, and handling of these materials. The occasional acute human poisoning cases are mainly due to accidents or suicidal misuse of these potent substances, but poisoning of wildlife is often caused by improper application or negligence as well as by unforeseen exposure of nontarget species to toxic doses of a natural pesticide or its residues. This chapter will focus on natural pest control agents of practical or historical importance. Some natural products, such as the avermectins, are treated elsewhere in this book, while the mode of action of some products, including that of pyrethrins and nicotine, is discussed elsewhere also in this book. Typically, chemical and common names are used throughout the chapter; commercial formulation types and product trade names can be found in current editions of various manuals (see, for example, Copping, 2004; Tomlin, 2003).

3.2  Insect control agents Some insects complete for our food and fiber, some damage construction materials, and some are important disease vectors in humans and animals. So, it is not surprising that there are so many natural insect control agents. Most of these were discovered by empirical screening of plants and, recently, other natural sources. Moreover, research on insect physiology and behavior made possible the commercialization of chemicals that can be used to manipulate insect development or behavior.

3.2.1  Botanical Insecticides 3.2.1.1  Pyrethrins (a)  Introduction Pyrethrum, also known as Dalmatian (or, inaccurately, Persian) insect powder, represents the dried flowers of the daisylike herbaceous perennial Tanacetum (Chrysanthemum) cinerariaefolium (Compositae) growing naturally along the east coast of the Adriatic Sea. It is the source of the economically most important botanical insecticides (Casida and Quistad, 1995a). The insecticidal properties of the flowers were discovered in Dalmatia, where the first commercial production of the plant began in 1840. The crop was later introduced in Japan, Kenya, Tanzania, China, Ecuador, New Guinea, Australia, Tasmania, and the United States. Since the early 1940s, the major producer of pyrethrum and pyrethrum extract has been Kenya providing about two-thirds of the world’s production. The insecticidal ingredients are called pyrethrins; they accumulate in the achenes of the flower heads and amount to 1–2% of the dried flower. The harvesting of flowers is labor intensive, but efforts to produce pyrethrins in tissue cultures on a large scale have not been fruitful (Jovetic´ and de Gooijer, 1995).

The key structural features of pyrethrins were recognized by Staudinger and Ruzicka in 1924 and the structures of all six insecticidal esters were clarified in the subsequent decades (reviewed by Crombie, 1995; Elliott and Janes, 1973; Matsui and Yamamoto, 1971). The exceptional biological properties and the lack of stability of the natural pyrethrins prompted many groups to find more stable synthetic analogs and these efforts eventually led to the development of numerous pyrethroid insecticides having a broad activity spectrum, excellent selectivity, and improved field stability (Elliott, 1996; Henrick, 1995; see elsewhere in this book). Mammalian toxicity studies as well as the experience of use for over a century demonstrate that pyrethrins are among the safest insecticides. (b)  Identity and Physicochemical Properties The insecticidal principles of the flower are six esters, collectively known as pyrethrins (Figure 3.1), formed by combination of two acids, chrysanthemic acid and pyrethric acid, and three alcohols, pyrethrolone, cinerolone, and jasmolone. The three esters of chrysanthemic acid and their relative amounts in a typical flower are pyrethrin I (38.0%), cinerin I (7.3%), and jasmolin I (4.0%). These are collectively known as pyrethrins I. The three esters of pyrethric acid and their relative amounts are pyrethrin II (35.0%), cinerin II (11.7%), and jasmolin II (4.0%). These are known as pyrethrins II. All six pyrethrins are high-boiling-point, viscous liquids and are soluble in hydrocarbons, alcohols, ethyl acetate, and halogenated solvents but poorly soluble in water. Note that the word “pyrethrin” was first used for what is now known as pellitorine, that is, N-(2-methylpropyl)(2E,4E)-2,4-decadienamide [18836-52-7], the main insecticidal ingredient of the roots of a North African plant, Anacyclus pyrethrum (Jacobson, 1971).

R' O

H

H

O H

O

R R pyrethrin I pyrethrin II cinerin I cinerin II jasmolin I jasmolin II allethrin

CH3 CO2CH3 CH3 CO2CH3 CH3 CO2CH3 CH3

R' CH=CH2 CH=CH2 CH3 CH3 CH2CH3 CH2CH3 H

Figure 3.1  Structures of natural pyrethrins and allethrin, a synthetic analog.

Chapter | 3  Pest Control Agents from Natural Products

Pyrethrin I �� IUPAC name: (Z)-(S)-2-methyl-4-oxo-3-(penta2,4-dienyl)cyclopent-2-enyl ()-trans-chrysanthemate. Chemical Abstract name: (1R,3R)-2,2-dimethyl-3-(2methyl-1-propenyl)cyclopropanecarboxylic acid (1S)-2methyl-4-oxo-3-(2Z)-2,4-pentadienyl-2-cyclopenten-1-yl ester. CAS Registry Number: [121-21-1]. Empirical formula: C21H28O3; molecular weight: 328.4. The solubility of pyrethrin I in water is 0.2 mg/l; the log P is 5.9. Pure pyrethrin I is levorotatory: [ ]20 D  14.0 (isooctane). Pyrethrin II �� IUPAC name: (Z)-(S)-2-methyl-4-oxo-3(penta-2,4-dienyl)cyclopent-2-enyl pyrethrate. ChemicalAbstractname:(1R,3R)-3-[(1E)-3-methoxy-2-methyl3-oxo-1-propenyl]-2,2-dimethylcyclopropanecarboxylic acid (1S)-2-methyl-4-oxo-3-(2Z)-2,4-pentadienyl-2cyclopenten-1-yl ester. CAS Registry Number: [121-29-9]. Empirical formula: C22H28O5; molecular weight: 372.4. The solubility in water is 9.0 mg/l, and the log P is 4.3. Pure pyrethrin II is dextrorotatory: [ ]20 D  14.7 (isooctane, ethyl ether). Cinerin I �� IUPAC name: (Z)-(S)-3-(but-2-enyl)-2-methyl4-oxocyclopent-2-enyl ()-trans-chrysanthemate. Chemical Abstract name: (1R,3R)-2,2-dimethyl-3-(2methyl-1-propenyl)cyclopropanecarboxylic acid (1S)3-(2Z)-2-butenyl-2-methyl-4-oxo-2-cyclopenten-1-yl ester. CAS Registry Number: [25402-06-6]. Empirical formula: C20H28O3; molecular weight: 316.4. Pure cinerin I is levorotatory: [ ]20 D  22.3 (hexane). Cinerin II �� IUPAC name: (Z)-(S)-3-(but-2-enyl)-2-methyl4-oxocyclopent-2-enyl pyrethrate. Chemical Abstract name: (1R,3R)-3-[(1E)-3-methoxy-2methyl-3-oxo-1-propenyl]-2,2-dimethylcyclopropanecarboxylic acid (1S)-3-(2Z)-2-butenyl-2-methyl-4-oxo-2cyclopenten-1-yl ester. CAS Registry Number: [121-20-0]. Empirical formula: C21H28O5; molecular weight: 360.4. Pure cinerin II is dextrorotatory: [ ]20 D  16.0 (isooctane). Jasmolin I �� IUPAC name: (Z)-(S)-2-methyl-4-oxo-3-(pent2-enyl)cyclopent-2-enyl ()-trans-chrysanthemate. Chemical Abstract name: (1R,3R)-2,2-dimethyl-3-(2methyl-1-propenyl)cyclopropanecarboxylic acid (1S)2-methyl-4-oxo-3-(2Z)-2-pentenyl-2-cyclopenten-1-yl ester. CAS Registry Number: [4466-14-2]. Empirical formula: C21H30O3; molecular weight: 330.4.

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Jasmolin II �� IUPAC name: (Z)-(S)-2-methyl-4-oxo-3-(pent2-enyl)cyclopent-2-enyl pyrethrate. Chemical Abstract name: (1R,3R)-3-[(1E)-3-methoxy-2methyl-3-oxo-1-propenyl]-2,2-dimethylcyclopropanecarboxylic acid (1S)-2-methyl-4-oxo-3-(2Z)-2-pentenyl-2cyclopenten-1-yl ester. CAS Registry Number: [1172-63-0]. Empirical formula: C22H30O5; molecular weight: 374.4. (c)  Stability of Pyrethrum Extract In the flower head, the pyrethrins are protected from photodecomposition (see, for example, Atkinson et al., 2004), and the insecticidal activity of the fresh, unground pyrethrum flowers is not reduced at temperatures up to 80°C. The pure components, pyrethrin I and pyrethrin II in particular, are readily oxidized and isomerized aerially and photochemically at ambient temperature (Bullivant and Pattenden, 1976; Otieno and Pattenden, 1980). The compounds are best stored in acetone at 25°C in the dark (Godin, 1968). In sunlight, pyrethrin mixture has a half-life of 10–12 min, which limits its outdoor uses. Burridge and Haya (1997) found that the half-life of cinerin I, used as a model pyrethrin component, was less than 6 h in seawater at 10°C. Sunscreens appear to protect the active ingredients from photodegradation (see, for example, Minello et al., 2005). As expected from laboratory studies, formulated pyrethrins are rapidly degraded under field conditions. For example, 1 h after treatment of pepper and tomato cultivations with a diatomaceous earth formulation containing 0.2% pyrethrins and 1.0% piperonyl butoxide (PB) synergist (application rate: 5.4 and 27.2 g/acre of pyrethrins and PB, respectively), the pyrethrin I  II residues on the pepper and tomato fruits were negligible (0.1 and 0.005 g/g, respectively) and the halflife of the insecticide was ca. 2 h for each crop (Antonious, 2004). The residual pyrethrin I content of peaches treated with 44 g/ha (recommended application rate) of a formulation containing 4% pyrethrins was 16 g/kg (detection limit) for pyrethrin I 1 day after treatment while at 220 g/ha treatment the half-life of this component was 2.6 days with residues 41 g/kg on the fruit 3 days after treatment (Angioni et al., 2005). The environmental fate of pyrethrins has been reviewed (Crosby, 1995; U.S. EPA, 2005a). (d)  Formulations and Uses The crude pyrethrum extract, called oleoresin, consists of the pyrethrins (25–30% total) accompanied by various resins, waxes, and pigments (Carlson, 1995). The refined commercial concentrate contains 45–60% pyrethrins, 20–25% light isoparaffins, 23–25% plant-derived triglycerides, terpenoids, and carotenoids, as well as 3–5% 2,6di-tert-butyl-4-methylphenol (BHT) as an added antioxidant (Maciver, 1995). Pyrethrins are commonly formulated as oil- or water-based aerosol sprays, emulsifiable concentrates, and dispersible or

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wettable powders. The commercial formulations typically contain 0.06–3.0% of the active ingredients often mixed with the noninsecticidal PB synergist, which inhibits oxidative detoxification thus enhancing the activity of the pyrethrins (Casida and Quistad, 1995b; Yamamoto, 1973). Products containing pyrethrins  PB on a solid support, such as diatomaceous earth (Korunic, 1998), are also marketed. Finely ground pyrethrum flowers are also used to make mosquito coils with 0.2–0.3% pyrethrum content that, when burning, both kill and repel mosquitoes. Pyrethrins are frequently mixed with other botanical or synthetic insecticides. For example, of the 1447 pyrethrin-containing products registered by the U.S. Environmental Protection Agency (EPA) in 2004, just 28 products formulated for end use contained pyrethrins as the only active ingredient. Pyrethrins are used in public health, on stored products, on domestic and farm animals, in aquaculture, as well as on ornamentals, greenhouse crops, fruits, vegetables and other field crops, and also for mothproofing textiles in museums. Pyrethrins are also available in shampoo to control ectoparasites such as scabies and lice in pet animals and humans (Anadón et al., 2009; Leone, 2007). Furthermore, pyrethrins are the active ingredients of disinfection sprays used in passenger aircrafts to prevent the spread of diseasecarrying insect vectors (see, for example, Berger-Preiß et al., 2004). PB-synergized pyrethrin formulations have also been applied aerially to large urban areas to control adult Culex spp. mosquitoes, the vectors of the human pathogen West Nile virus, without any human health or ecotoxicological consequences (Carney et al., 2008; Gammon, 2007; Schleier et al., 2008). Pyrethrum extracts are also used in aquaculture as brief bath treatments to control salmon lice (Boxaspen and Holm, 2001). The different uses of pyrethrum have been reviewed (Gerberg, 1995; Kennedy and Hamilton, 1995; Silcox and Roth, 1995). (e)  Biological Properties Mode of action  The mode of action of pyrethrins can be inferred from investigations that used synthetic pyrethroids instead of the natural mixture. Of the synthetics, allethrin [584-79-2] (Figure 3.1) is the closest structural relative of the natural products. Although some of the newer analogs possess specific structure-dependent properties not shared by the natural products, the information obtained for pyrethroids is instructive in understanding the mode of action of the pyrethrins as well. The subject has been extensively reviewed (Bloomquist, 1996; Soderlund, 1995; Soderlund et al., 2002) and is discussed in depth elsewhere in this book. Electrophysiological experiments with nerve preparations from both invertebrates and vertebrates revealed that the principal target sites of pyrethrins and pyrethroids are voltage-sensitive Na channels. These ion channels are involved in the propagation of action potentials along the nerve membrane and in the regulation of neurotransmitter

Hayes’ Handbook of Pesticide Toxicology

discharge from presynaptic sites. The lipophilic pyrethrins selectively bind to open (activated) Na channels and slow or delay their closing (inactivation), thus allowing the influx of sodium ions. The continuous inward ion current causes depolarization and repetitive firing of the nerve membrane that are responsible for the various pharmacological effects of these neurotoxins. Insect sodium channels are more sensitive to pyrethroids than mammalian sodium channels, providing partial explanation to the selective toxicity of these substances. The first primary structure determined for a Na channel protein was from the eel electroplax (Noda et al., 1984). The 1820 amino acid residues are grouped into four internally homologous domains, each containing six -helical transmembrane segments. Sodium channel proteins and their genes have been sequenced and characterized from insects and mammals, including human brain (Ahmed et al., 1992), are structurally related, and the operation of the channel is now well understood at the molecular level (Catterall et al., 2007; Strong et al., 1993; Wang and Wang, 2003). Electrophysiological experiments and radioligandbinding assays as well as molecular modeling studies with wild-type and mutated Na channel proteins have provided evidence for a specific pyrethroid binding site, but the isolation of the membrane-embedded receptor remains elusive. Two types of insecticidal action of pyrethrins can be distinguished: rapid paralyzing knockdown that is characteristic to pyrethrin I and slowly developing kill which is a typical feature of pyrethrin II (Briggs et al., 1974; Sawicki and Thain, 1962). In acute poisoning, the pyrethrins and most noncyanohydrin ester pyrethroids affect mainly the peripheral nervous system and produce the following symptoms in mammals: hyperexcitation, prostration, and wholebody tremors with clonic and, occasionally, tonic seizures before death at lethal doses. This syndrome, resembling that produced by DDT (dichlorodiphenyltrichloroethane), was designated as the T syndrome (Verschoyle and Aldridge, 1980) or Type I syndrome (Gammon et al., 1981; Lawrence and Casida, 1982). Pyrethroids of the cyanohydrin ester type act primarily on the central nervous system (CNS), and the typical symptoms are profuse salivation, pawing and burrowing behavior, and choreoathetosis (sinuous writhing of the body). This particular set of symptoms was designated as the CS syndrome or Type II syndrome. For Type I pyrethroids, electrophysiology indicates repetitive firing due to a transiently modified open channel. By contrast, Type II compounds persistently retain the Na channel in an open state and block the action potential without repetitive firing (Lund and Narahashi, 1983). Pyrethroids, although often at high micromolar concentrations, target additional biochemical and pharmacological processes, including the Ca2-stimulated ATPase activity in squid nerves (Clark and Matsumura, 1982; Grosman and Diel, 2005), voltage-gated calcium ion channels (reviewed by Shafer and Meyer, 2004), the nicotinic acetylcholine

Chapter | 3  Pest Control Agents from Natural Products

receptor (nAChR) of the electric eel (Abbassy et al., 1983), -aminobutyric acid (GABA) receptor–ionophore complex in rat brain synaptic membranes (Lawrence and Casida, 1983), norepinephrine release and Ca2 uptake in rat brain synaptosomes (Doherty et al., 1987), protein phosphorylation in rat brain synaptosomes (Enan and Matsumura, 1991), the voltage-gated chloride channel in rodent skeletal muscle and neuronal fibres (Forshaw et al., 2000), and the glutamatergic system in mouse cortical and spinal cord tissues (Shafer et al., 2008). These effects, often observed for other types of insecticides as well, are probably not responsible for the primary action of pyrethroids but they could be implicated in some neurotoxic effects observed in mammals for these compounds (reviewed by Ray and Fry, 2006; Soderlund et al., 2002). Metabolism and excretion  Both insects and mammals metabolize pyrethrins rapidly, but in insects, penetration into the CNS and other target tissues via the tracheal system is relatively fast (Gerolt, 1975), which contributes significantly to the selective toxicity of these insecticides. The principal detoxification process of pyrethrins in insects and rodents is oxidation at multiple sites, and the hydrolysis of the ester group(s) is generally not important (Casida et al., 1971; Class et al., 1990; Elliott et al., 1972; reviewed by Casida and Quistad, 1995b). Recent studies (see, for example, Price et al., 2008) indicate that high doses of pyrethrum extract can induce mammalian hepatic cytochrome P-450 (CYP) oxidases. Preadministration of the CYP oxidase inhibitor PB increases substantially the toxicity of pyrethrins in insects (Ando et al., 1983; Yamamoto, 1973) but much less so in mammals. In fact, the initial inhibition of oxidative metabolism by PB 3–12 h after intraperitoneal administration to mouse is followed by induction of microsomal oxidase activity (Škrinjaricˇ-Špoljar et al., 1971; see also Bond et al., 1973; Springfield et al., 1973). Casida and Quistad (1995b) reviewed the available mammalian excretion studies on the metabolites of pyrethrins I and II. The major urinary metabolites of radiolabeled pyrethrin I and pyrethrin II were dihydrodiols, resulting from the pentadienyl side chain via epoxidation and subsequent hydration; alcohol, aldehyde, and carboxylic acid modifications from regioselective oxidation of the isobutenyl side chain of pyrethrin I; and carboxylic acid from hydrolysis of the methoxycarbonyl group of pyrethrin II. These metabolites are presumably excreted as glucuronidated derivatives by analogy to a study on allethrin (Class et al., 1990). The metabolism of pyrethrins in humans has not been investigated in detail but it is assumed to proceed similarly to that in rodents. Recent human studies demonstrated that the common urinary metabolite trans-chrysanthemumdicarboxylic acid (CDCA) is excreted with a ca. 4 h half-life after oral intake of 3.3–5.4 g of pyrethrin I per kg body weight (Leng et al., 2006). In the same laboratory study, after a 1-day inhalation exposure of pyrethrins, the urinary concentration

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of CDCA ranged from 1 to 54 g/l. Furthermore, urine analysis for CDCA of pilots who had mixed, loaded and applied an ultra-low-volume pyrethrum formulation (1.25 h total exposure time), indicated that the highest daily pyrethrinexposure was 0.03 g/kg, which is one-thousandth of the acceptable daily intake (Gerry et al., 2005). (f)  Toxicity to Test Animals The pyrethrum extract contains not only the six cyclopropanecarboxylic acid esters but, in varying amounts, additional components that contribute to the biological activity of the pyrethrins. Thus, early toxicity data obtained by different laboratories for unpurified oleoresins with 30–35% pyrethrin content varied greatly. In 1944, enrichment methods providing concentrates of 77.8% pyrethrin content became available (Barthel et al., 1944), permitting more reproducible toxicity assays (Malone and Brown, 1968). Subsequently, the toxicity of each single component was determined. Acute and chronic toxicity data of pyrethrins were compiled in several reviews (Barthel, 1973; Griffin, 1973; Negherbon, 1959; Pillmore, 1973). Additional aspects of pyrethrin toxicology were discussed recently (Litchfield, 1985; Ray, 1991; U.S. EPA, 2005a). Acute toxicity  Representative acute toxicity data for pyrethrum extracts of various purity as well as data for some purified components are given in Table 3.1. It must be pointed out that pyrethrins are highly toxic to fish and some aquatic insects and crustaceans on which fish feed (Bridges and Cope, 1965; Mauck et al., 1976). Pyrethrins are toxic to bees with a contact LD50  22 ng per honeybee for a 57.6% pure extract (see Gabriel and Mark, 1995). In the field, however, the repellency of the insecticide and proper timing of the application mitigate the risk. Pyrethrins and their synthetic analogs are more insecticidal at low than at high temperatures (Gammon, 1978; Wang et al., 1972), thus application of pyrethrins at night could provide better control. Because there is a 10°C difference in the body temperatures of mammals and insects, this phenomenon contributes to the selective toxicity of these compounds. However, this negative temperature coefficient was also observed in lizards (Talent, 2005) and in frogs (Cole and Casida, 1983; van den Bercken, 1977) raising some ecotoxicological concern. The toxicity of pyrethrum extract to fish is higher in acidic (pH 6.5) than in alkaline (pH 8.5) water and increasing water hardness slightly enhances fish toxicity but these changes are most likely related to chemical degradation (Mauck et al., 1976). Subchronic and chronic toxicity  Due to their widespread indoor uses (on animals, on stored products, in the household, etc.), exposures to pyrethrins could be quite frequent and/or lasting, necessitating long-term studies.

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Table 3.1  Acute Toxicity of Pyrethrins Animal

Assay

LD50 (mg/kg) or LC50 (mg/l)

Notes, other data

References

Rat

oral

820

Crude oleoresin

Carpenter et al. (1950)

Rat

oral

1870

Purified extract

Carpenter et al. (1950)

Rat, female

oral

584

20.9% pure pale extract

Malone and Brown (1968)

a

Rat, female

oral

715

77.8% pure

Malone and Brown (1968)

Rat, male

oral

260–420

Pyrethrin I

Casida et al. (1971)

Rat, male

oral

600

Pyrethrin II

Casida et al. (1971)

Rat, male

oral

470

74% pure extract

Barthel (1973)

Rat, female

oral

263

74% pure extract

Barthel (1973)

Rat

oral

1440

20% solution

Bond et al. (1973)

Rat, male Rat, female Rat, female

oral oral

Rat

inhalation

Mouse, female

oral

Guinea pig Rabbit Rabbit Dog

57.6% pure

d

ic

oral ip

Malone and Brown (1968)

Subchronic LC50  3.4 mg/l air; 58% pure extract

Schoenig (1995); U.S. EPA (2005a)

285

77.8% purea

Malone and Brown (1968)

2500

e

Lawrence and Casida (1982)

1500 102

dermal dermal

19,800 2000

f

Oleoresin

Shimkin and Anderson (1936)

f

Oleoresin

Shimkin and Anderson (1936) a

77.8% pure

c

iv

a

oral

2000

57.6% pure

Catfish Bluegill

Negherbon (1959)

LD  6–8 mg/kg g

Bobwhite quail

Catfish

Schoenig (1995)

57.6% pure

77.8% pure

Steelhead trout

Malone and Brown (1968)

b

1565

Rainbow trout

Verschoyle and Barnes (1972)

77.8% pure

1350

pv

Frog

Schoenig (1995)

a

Chicken

Mallard duck

Schoenig (1995)

b

LD 5 mg/kg

iv dermal

Guinea pig

1030

57.6% pure

c

Rat

Mouse, male

2370

b

Malone and Brown (1968) Gabriel and Mark (1995) h

5-day

LC50  5620 ppm

Gabriel and Mark (1995)

sc

e

Cole and Casida (1983)

48-h 96-h 48-h 96-h 48-h

5.8

LC50  54 ppb

i

Bridges and Cope (1965)

LC50  25 ppb

j

Mauck et al. (1976)

LC50  82 ppb

i

LC50  114 ppb LC50  74 ppb

Bridges and Cope (1965) j

Mauck et al. (1976)

i

Bridges and Cope (1965)

j

Bluegill

96-h

LC50  41 ppb soft water, pH 6.5

Mauck et al. (1976)

Bluegill

96-h

LC50  87 ppbj soft water, pH 9.5

Mauck et al. (1976)

Bluegill

96-h

LC50  46 ppbj very hard water, pH 8.2

Mauck et al. (1976)

Coho salmon

96-h

LC50  39 ppbi

Pillmore (1973)

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125

Table 3.1  (Continued) Animal

Assay

Coho salmon

96-h

American lobster

48-h

LD50 (mg/kg) or LC50 (mg/l)

Notes, other data

References

LC50  39 ppbj

Mauck et al. (1976)

LC50  0.73 ppb

k

i

Burridge and Haya (1997)

Stonefly

48-h

LC50  6.4 ppb

Bridges and Cope (1965)

Daphnia pulex

48-h

LC50  25 ppbi

Pillmore (1973)

a

Refined nitromethane concentrate of pyrethrum extract. The ratio of pyrethrins I to pyrethrins II of the sample was 1.85. c Approximate lethal dose of pyrethrins. d Intracerebral injection. e Mixture of 40% pyrethrins I and 46% pyrethrins II. f Recalculated from original mortality data. Oleoresin contained 14% pyrethrins I and II. g Perivisceral injection. h Dietary toxicity; lethargy and reduced body weight but no mortality were observed. i Static tests with extract containing 24.6% pyrethrins. j Static tests at 12°C with extract containing 20% pyrethrins. k For the most sensitive, stage IV larvae. Twenty-five percent pyrethrum extract with PB synergist was used. b

Based on observations by Kimbrough et al. (1968) that pyrethrum, especially synergized pyrethrum, produced liver enlargements and cellular effects, including margination and cytoplasmic inclusions in rats, Springfield et al. (1973) examined the related biochemical changes. Oral administration of a 20% pyrethrin mixture to male rats at 200 mg/kg for 23 days resulted in significantly increased total lipid concentration as well as liver enlargement that was independent from the former. Nicotinamide adenine dinucleotide phosphate (NADPH)–cytochrome c reductase activity and CYP concentration were also increased with a concomitant elevation in drug metabolic activities. Hepatic protein and water contents were not different from controls. The observed changes reversed to normal on cessation of treatment. A more recent 42-week study examined the tumorinducing effects in rats of a 57% pyrethrin extract in the diet (Finch et al., 2006; Price et al., 2007). The no-effect level for both liver and thyroid tumors was 100 ppm pyrethrins in both sexes. Significant increases were observed in liver and thyroid gland weights of male rats given 8000 ppm pyrethrins and female rats given 3000 and 8000 ppm pyrethrins for 14 and 42 days. Treatments of male rats with 8000 ppm pyrethrins and female rats with 3000 and 8000 ppm pyrethrins resulted in significant increases in hepatic microsomal CYP content and some CYP-dependent activities. On cessation of treatment all the observed effects were reversible in both sexes. Moreover, the effects seen with pyrethrins treatment were similar to those produced by subchronic doses of phenobarbital, a known hepatic metabolism inducer and a known liver and thyroid gland tumor promoter in rodents. Earlier epidemiological studies, however, revealed no liver or thyroid gland tumor risks in human subjects receiving phenobarbital for years at doses comparable to those producing carcinogenesis

in rodents, so pyrethrins are not expected to produce such tumors in humans at normal level of exposure. Results of subchronic mammalian toxicology studies with pyrethrum extract containing 57.6% pyrethrins were summarized by Schoenig (1995). In studies with rodents, characteristic pyrethrin poisoning symptoms (labored respiration, tremors, hyperactivity, and death) were observed at dietary concentrations of 10,000 ppm or higher. No treatment-related effects were seen at 300 and 1000 ppm for mice and rats, respectively. In a 104-week chronic feeding study with rats, the highest concentration without treatment-related effects was 100 ppm. For both sexes, the 3000-ppm concentration resulted in small increases in hyperplasia and follicular cell adenoma. These symptoms were attributed to lowered thyroid hormone level and elevated thyroid stimulatory hormone level caused by intense liver metabolic processes provoked by the extreme pyrethrum doses. A small increase in the incidence of keratoacanthomas of the skin of male rats was also observed. In an 18-month mouse feeding study, no treatment-related effects were seen below 100 ppm. In a 1-year study with beagle dogs, no toxic effects were found for concentrations up to 500 ppm pyrethrins in the diet. Anemia, increased serum glutamic pyruvic transaminase levels, and increased liver weight were observed at 2500 ppm. Irritation and inhalation  Reviewing early works with rats, guinea pigs and rabbits, Barthel (1973) concluded that the dermal toxicity of pyrethrins is negligible with LD50 values ranging from 1350 to about 5000 mg/kg. Schoenig (1995) reported no ill effects in a 21-day rabbit dermal toxicity assay at 1000 mg/kg dose. The most detailed inhalation toxicity study on pyrethrins was done with rats by Carpenter et al. (1950). With

126

27 to 85 30-min exposures to 17.6 mg/m3 aerosolized pyrethrins, pathology showed no treatment-related effects, and with several thousandfold higher extreme doses, minor lung congestion was induced only. In recent rat inhalation toxicity studies (Schoenig, 1995; see also U.S. EPA, 2005a), a 4-h exposure to the liquid aerosol of pyrethrins gave an LC50 of 3400 mg/m3. In the 90-day test, no systemic toxicity was seen up to 11 mg/m3, but at doses higher than 30 mg/m3 anemia was evident. At 356 mg/m3, first labored breathing, hyperactivity, and tremor were apparent and pathology revealed increased liver weights as well as microscopic changes in the respiratory tract indicative of irritation. Teratogenic, carcinogenic, and mutagenic effects Williams (1973) reviewed the results of early mammalian toxicity studies. Briefly, there was no evidence of teratogenic effects in a one-generation reproduction study with female rats fed 5000 ppm pyrethrins in the diet. No adverse effects were seen in a teratogenicity study with female rabbits given 90 mg/kg orally for 8 days during gestation. In a rat teratogenicity study with a technical grade pyrethrum extract containing 10.94% pyrethrin I and 9.06% pyrethrin II, minor, but dose-dependent increases in the occurrence of resorptions as well as reductions in the average proportions of live fetuses were noted at 50–150 mg/ kg daily oral doses from the 6th to the 15th days of gestation (Khera et al., 1982). Results of recent teratology studies meeting current data requirements were reviewed by Schoenig (1995; see also California EPA, 1996; U.S. EPA, 2005a). Upon oral application of pyrethrum extract containing 57.6% pyrethrins to female rats on gestation days 6–15, maternal toxicity (convulsions, tremors, and mortality) was observed at daily doses greater than 150 mg/ kg, but no developmental toxicity was seen at doses up to 600 mg/kg/day. The two-generation reproduction study with pyrethrins up to 3000 ppm in the diet gave no evidence of parental toxicity of the F0 generation. Neonatal toxicity in the form of decreased pup body weights in both the F0 and the F1 generations was observed at 1000 and 3000 ppm. The highest concentration without treatmentrelated reproductive toxicity was 100 ppm in this animal. In a similar series of experiments, rabbits proved to be more vulnerable and high postimplantation loss was noted at 600 mg/kg/day but no treatment-related maternal toxicity was observed at 25 mg/kg/day, the lowest dose used. This study did not indicate developmental toxicity. No oncogenic effects were seen in rodent studies at 100 ppm dietary concentrations of pyrethrins (California EPA, 1996; Schoenig, 1995; U.S. EPA, 2005a). The induction of liver and thyroid gland tumors by high doses of dietary pyrethrins has been discussed earlier (see above and Finch et al., 2006; Price et al., 2007; see also U.S. EPA, 2004). Pyrethrum extract was not mutagenic in the Ames test, no increases in chromosomal aberrations were seen in

Hayes’ Handbook of Pesticide Toxicology

Chinese hamster ovary cells, and no increase in unscheduled DNA synthesis was observed (California EPA, 1996; Schoenig, 1995; see also Ashwood-Smith et al., 1972; Moriya et al., 1983; U.S. EPA, 2005a). Endocrine effects  The binding of pyrethrins and their synthetic analogs to various steroid hormone receptors was recently examined. In a human skin fibroblast androgen receptor assay, a 20% pyrethrum extract inhibited methyltrienolone binding with a Ki of 1.5  105 M and also inhibited testosterone binding to sex hormone binding globulin at 104 M (Eil and Nisula, 1990). According to Soto et al. (1995), pyrethrum extract lacks estrogenic activity. Among a series of pyrethroids tested in T47D human breast cancer cell line, 30 M d-trans-allethrin antagonized the progestagenic effect of 5 nM progesterone (Garey and Wolff, 1998). Treatment  Because pyrethroids, including pyrethrins, affect multiple neurotransmitter systems, treatment of systemic poisoning poses a challenge. The toxic symptoms caused by intravenous administration of permethrin at 40 mg/kg (ED95) in mice could be partially alleviated by intraperitoneal pretreatment with diazepam (10 mg/kg), aminooxyacetic acid (50 mg/kg), or cycloheximide (1 mg/ kg) (Staatz et al., 1982). The anticonvulsant diazepam at a 3 mg/kg intraperitoneal dose was more specific and potent than phenobarbital in protecting mice from LD95 doses of Type II pyrethroids and increased the intracerebroventricular LD50 values of deltamethrin or permethrin six- to ninefold (Gammon et al., 1982). Mephenesin prevented choreoathetosis at moderate doses but gave full protection against both types of pyrethroids only at high doses that also caused profound muscle relaxation (Bradbury et al., 1983). Methocarbamol, alone or in combination with atropine, which alone blocked salivation only, reduced poisoning symptoms and mortality in pyrethroidtreated rats (Hiromori et al., 1986). Ray and Fry (2006) have recently summarized the potential treatment options. (g)  Toxicity to Humans Systemic poisoning incidents  Acute pyrethrum poisoning cases are extremely rare. In fact, pyrethrum preparations were once recommended as internal anthelmintic agents (McLellan, 1964). According to recent pyrethrinpoisoning incidents reports (U.S. EPA, 2005b), the majority of cases were related to getting pyrethrin-based lice shampoo in the eyes of children causing eye irritation, corneal abrasion, lacrimation, eye burns, and blurred vision. Additional reported symptoms resulting from various acute exposures are dyspnea (difficult breathing), coughing and bronchospasm, nausea and vomiting, as well as dermal effects (see below and also Paton and Walker, 1988). Chronically exposed persons may present hypersensitivity pneumonitis characterized by chest pain, cough, dyspnea, and bronchospasms.

Chapter | 3  Pest Control Agents from Natural Products

Accidental ingestions by children, including a case of a 2-year-old girl who died after eating approximately 14 g of pyrethrum powder, were reported in the late 19th century (see Ray, 1991). Repeated dermal (and possibly oral) exposure of a 2-year-old girl to an over-the-counter pyrethrins  PB pediculocide caused stuttering and incoordination, symptoms consistent with peripheral neurotoxic effects of the insecticide, that resolved without treatment (Hammond and Leikin, 2008). Wax and Hoffman (1994) reported a fatality associated with the inhalational and dermal exposure to a pet shampoo containing pyrethrins (0.06%) and PB (0.6%). Death was attributed to irreversible bronchospasm; postmortem findings (thickened basement membranes, goblet cell hyperplasia, prominent mucous gland, mucous plugging, and smooth muscle hypertrophy as well as eosinophilic infiltrates) were consistent with the natural history of fatal asthma. A similar poisoning was reported for an 11-year-old asthmatic girl who was bathing a dog with a pet shampoo containing 0.2% pyrethrins when severe shortness of breath with wheezing developed. In spite of aggressive hospital treatment including intubation, she died about 2.5 h after the initial exposure (Wagner, 2000). Irritation and sensitization  Upon dermal contact, even very low doses of pyrethrins and pyrethroids can produce local paresthesia (a sensation of tingling, burning, stinging, numbness, and itching of the exposed skin). Such transient paresthesia, probably due to repetitive firing of skin nerve terminals, is typically encountered by individuals occupationally exposed to these substances (but see Paton and Walker, 1988). Allergy from inhalation of or direct dermal contact with pyrethrum flowers during harvesting and processing or with unrefined pyrethrum extracts is not uncommon (Barthel, 1973; Garcia-Bravo et al., 1995; McCord et al., 1921). The initial symptoms of pyrethrum dermatitis are mild erythema covering hands and face which, on further contact, can develop into edema and blistering (see Rickett et al., 1972). Understandably, the allergenic properties of the flower, its constituents, and the synthetic pyrethroids were thoroughly investigated (Taplin and Meinking, 1990). Testing immediate and delayed hypersensitivity reactions in guinea pigs previously sensitized to pyrethrum, Rickett and Tyszkiewicz (1973) identified the major allergens in the 0.9% saline extract of pyrethrum flowers. The allergenic agents were tentatively assigned as 60–200 kDa glycoproteins. The refined pyrethrum extract and pyrethrin II were not allergenic. Crude oleoresin contained traces of another allergen, pyrethrosin, a sesquiterpene lactone possessing a reactive -methylene moiety (Mitchell et al., 1972; Rickett and Tyszkiewicz, 1973). Sensitization to pyrethrum frequently occurs with individuals sensitive to the pollen of Ambrosia spp. (ragweed) (Feinberg, 1934; see also Carlson and Villaveces, 1977). Pyrethrosin-related lactones are the major allergens of

127

noninsecticidal ornamental chrysanthemum species (Hausen and Schulz, 1973). A comprehensive health risk assessment of pyrethrins is available (U.S. EPA, 2005a). Treatment  Because pyrethrins are readily metabolized and excreted, treatment of pyrethrin poisoning is mainly symptomatic and supportive. Pneumonitis, resulting from aspiration of kerosene or other hydrocarbons used in the insecticide formulation, may complicate poisoning incidents. Because of the potential irritancy of pyrethrins, proper decontamination following exposure is important. Dermatitis and allergic reactions caused by occupational exposure on pyrethrum-growing farms and in production facilities are preventable by minimizing exposure to the irritant (Gnadinger, 1945; Moore, 1975). Pyrethrin spray-related hypersensitivity pneumonitis could be treated with prednisone and ampicillin (Carlson and Villaveces, 1977). Tucker et al. (1983) recommended vitamin E (-tocopherol) oil for the immediate local treatment of the cutaneous sensation or paresthesia among individuals dermally exposed to synthetic pyrethroids. The mode of action of vitamin E appears to be the selective block of pyrethroid-modified Na channels as was shown in situ in rat cells (Song and Narahashi, 1995).

3.2.1.2  Nicotine (a) Introduction Nicotine, a structurally simple alkaloid (Figure 3.2), is a most notorious botanical insecticide. It is the main bioactive component of the tobacco plants Nicotiana tabacum, N. glauca, and N. rustica (Solanaceae) and thought to function as a herbivore repellent. Nicotine is also found in the leaves of the Australian shrub Duboisia hopwoodii (Solanaceae), which is used by the Aborigines as a stimulant and hunting aid. It is also present in a number of other plants of the families of Lycopodiaceae, Crassulaceae, Leguminosae, Chenopodiaceae, and Compositae (Leete, 1983). It is noteworthy that nicotine is

H

N

H

H

1' 5' N

N H N

N

(S)-nicotine

N

nornicotine

cotinine

H

H

N

myosmine

N H

N H

N

N

N

anabasine

Figure 3.2  Structures of tobacco alkaloids.

anatabine

O

128

also a natural constituent of several Solanaceous food plants, including potatoes, tomatoes, eggplant as well as pepper, and can be detected at g/kg level in the fruits of these plants (reviewed by Andersson et al., 2003). Nicotine is biosynthesized in the roots and thereafter translocated to the aerial parts, reaching, on a dry-weight basis, 1–8% content in the leaves of N. tabacum and 2–18% in N. rustica. The pure alkaloid was first isolated by Posselt and Reimann in 1828 and the structure was determined by Pinner in 1893. (b)  Identity, Physicochemical Properties, and Uses IUPAC name: (S)-3-(1-methylpyrrolidin-2-yl)pyridine. Chemical Abstract name: 3-[(2S)-(1-methyl-2-pyrrolidinyl)] pyridine. CAS Registry Number: (S)-nicotine: [54-11-5]. CAS Registry Numbers for other tobacco alkaloids: (R)-nicotine [25162-00-9]; racemic nicotine [2208374-5]; nicotine sulfate [65-30-5]; (S)-anatabine [58149-7]; (S)-cotinine [486-56-6]; myosmine [532-12-7]; (S)-nornicotine [494-97-3]. Empirical formula: C10H14N2; molecular weight: 162.2. Physicochemical properties of nicotine  Pure nicotine is a colorless liquid that boils at 246–247°C; its freezing point is below 79°C. Its density is 1.009 g/cm3 at 20°C. The free base is fairly volatile with a vapor pressure of 4.25  102 mmHg. The concentration of nicotine in the vapor phase, as determined by the “air bubbling method,” is about 28 ppm at 25°C (see Jackson, 1941). Nicotine is hygroscopic and freely miscible with water, ethanol, ethyl ether, and most organic solvents. Its pKa1  3.09 and pKa2  8.18. The log P value of the nonionized alkaloid is 0.93 (Chamberlain et al., 1996); according to Seckar et al. (2008), the log P values for nicotine hemisulfate at pH  3, 7 and 9 are 1.60, 0.16 and 1.25, respectively. Pure natural nicotine is levorotatory: [ ]20 D  163.9 . Hereinafter, “nicotine” will generally refer to the natural product, that is, the (S)-isomer. Stability  Upon exposure to air pure nicotine turns brown with a characteristic odor reminiscent of tobacco. Its watersoluble salts are more stable. At 30°C and in the presence of oxygen, the photooxidative degradates of the alkaloid include cotinine, myosmine, nicotinic acid, methylamine, and ammonia. In the field, however, the persistence of nicotine is sufficiently long for insecticidal purposes. For example, nicotine sprayed on mustard greens had a half-life of 4.5 days at 10–50 ppm treatments, indicating that the alkaloid penetrated into the wax layer of the leaf and thus was protected from aerial dissipation (Gunther et al., 1959). A vapor photolysis pilot study by Seckar et al. (2008) suggested that the half-life of nicotine is approximately 12 h, or equivalent to 1 day of natural sunlight. Due to the high global prevalence of tobacco smoking, nicotine and its primary stable metabolite, cotinine, are

Hayes’ Handbook of Pesticide Toxicology

ubiquitous contaminants in surface waters and wastewater at as high as ppb (g/l) concentrations (see, for example, Buerge et al., 2008; Huerta-Fontela et al., 2007). Formulations and agricultural uses  The tobacco plant was introduced to Europe in 1559 from the Americas where it had long been cultivated by the American Indians primarily for smoking. From 1690 in Europe, tobacco dust and extracts were used to repel or kill insects and tobacco smoke was also used for fumigation. Tobacco was (re)introduced as an insecticide in the United States in 1814. Nowadays, nicotine is used in agriculture only on a limited scale, for example, in greenhouses and in organic farming. It may be noted that tobacco plant preparations are still used as affordable popular folk remedies so animal and human poisoning cases due to nicotine and related alkaloids are not uncommon. Nicotine can also be used to capture and restrain dangerous, unmanageable, or wild animals. This method relies on a projectile-type syringe loaded with nicotine solution and fired from a special CO2powered rifle. Cattle weighing up to 450 kg could thus be quickly immobilized with up to 3 mg/kg doses (Hayes et al., 1959) but the high amount of nicotine that should be handled represents a risk to users. Nicotine dust formulations are also available as dog and rabbit repellents. Nicotine is mainly obtained as a by-product of cigarette manufacturing. The commercial insecticide may contain traces of accompanying tobacco alkaloids. Nicotine has a systemic action and is used on fruits, vegetables, and ornamentals against a wide range of insects, including aphids, thrips, and whiteflies. It also has anthelmintic and molluscicidal properties. Nicotine and related insecticides have been reviewed (Schmeltz, 1971; Ujváry, 1999). Nicotine free base is commercialized as a concentrated, for example 40%, aqueous solution or fumigant formulation of 0.05–4.0% alkaloid content. Nicotine sulfate is sold as a dispersible powder and as a 40% aqueous solution (Black Leaf 40; since 1992 it is not registered in the United States). Nicotine is formulated also on an inert material support such as bentonite. Recently, emulsions containing nicotine and long chain fatty acids have been developed (Casanova et al., 2005). (c)  Biological Properties Nicotine is the main psychoactive component of tobacco, which is regularly smoked by hundreds of millions of people worldwide, and its biochemistry, pharmacology, and toxicology have been thoroughly investigated. It was, however, recently noted that toxicology data meeting current pesticide registration requirements are limited (U.S. EPA, 2008a). Early works on the various biological activities of nicotine are summarized by Larson et al. (1961). For a modern pharmacological treatment, the reader is referred to Taylor (2006) as well as to a review by Benowitz (1996). In this section, the acute biological properties of nicotine will be emphasized.

Chapter | 3  Pest Control Agents from Natural Products

Mode of action  By mimicking the excitatory neurotransmitter acetylcholine (ACh), nicotine exerts its pharmacological and neurotoxic effect in animals and humans by binding to a subset of cholinergic receptors, the nAChRs. The structural and functional diversity of the several nAChR subtypes known are better characterized for vertebrates than for insects (reviewed by Millar and Denholm, 2007; Tomizawa and Casida, 2005). In mammals, these ligandgated ion channel receptors are formed from pentameric arrangements of different homologous peptide subunits of which at least 17 are known. Receptors composed of different subunit combinations have distinct physiological and pharmacological properties and are differentially distributed at the neuromuscular junction and within the central and peripheral nervous system (see Gotti et al., 2007). The activation of nAChRs by nicotine and related agonists causes a rapid increase in cellular permeability to Na and Ca2, leading to depolarization and excitation that generate the release of various neurotransmitters. Prolonged application of nicotine or other agonists results in desensitization of the cholinergic receptor site and a lasting blockade. The effects of nicotine on nAChRs are not antagonized by atropine but can be selectively blocked by other agents (e.g. tubocurarine or -bungarotoxin). Nicotine can also act at noncholinergic sites. In insects, nAChRs were detected only in the central nervous system, both pre- and postsynaptically (Eldefrawi and Eldefrawi, 1997). The binding of nicotine is stereoselective. (S)-Nicotine was 10- to 60-fold more active than the nonnatural (R)enantiomer in vitro in some (Barlow and Hamilton, 1965; Romano and Goldstein, 1980; Zhang and Nordberg, 1993) but not in other (Ikushima et al., 1982) bioassays. The acute toxicities of the isomeric alkaloids also differ; for example, the intravenous LD50 value of the (S) and (R) stereoisomer in mice was 0.38 mg/kg and 2.75 mg/kg, respectively (Aceto et al., 1979). The recent discovery of a new class of insecticides, the neonicotinoids, has rekindled interest in nAChRs (Jeschke and Nauen, 2008; Yamamoto and Casida, 1999; see also elsewhere in this book). Absorption, metabolism, and excretion  The mammalian metabolism and pharmacokinetics of nicotine is well understood (Hukkanen et al., 2005). Nicotine base is readily absorbed through the skin, the mucous membranes, and, when inhaled, the lungs. Absorption is less from acidic solutions, rendering the commercial sulfate salt safer on dermal contact (Faulkner, 1933). Nicotine is not readily absorbed from the stomach unless intragastric pH is raised but intestinal absorption is far more efficient. Once absorbed, the alkaloid easily crosses the placenta and is also distributed in breast milk. Up to 90% of the absorbed nicotine is rapidly metabolized in the liver, the kidneys, and the lungs. The metabolites and any unaltered nicotine are eliminated by the urine. The

129

elimination half-life of nicotine in humans is 100–250 min after intravenous administration (see Hukkanen et al., 2005). Both in insects and in mammals, the primary metabolite of nicotine is cotinine (Figure 3.2) formed in a two-step oxidative process involving microsomal CYP and cytosolic aldehyde oxidase enzymes. Cotinine is also a trace tobacco constituent. It is a poor insecticide (Yamamoto, 1965) and inactive at the mammalian nAChR (Benowitz, 1995). Nevertheless, cotinine influences neurotransmitter release in the brain and affects the cardiovascular system and a number of enzymes. Cotinine levels are at least 10-fold higher than those of nicotine and its half-life is 4–8 times longer than that of the parent alkaloid; thus, its contribution to the overall pharmacology and toxicity of nicotine cannot be discounted (see Benowitz, 1996; Vainio et al., 1998). Cotinine is also the major metabolite in crops treated with nicotine insecticide (Gunther et al., 1959) and an ubiquitous contaminant of urban wastewater (see, for example, Gagné et al., 2006). Nornicotine, formed by oxidative demethylation, is another metabolite of nicotine. It is also present in Solanaceae usually as a mixture of the (S) or (R) stereoisomers, either of which may predominate. In studies with rat brain nAChRs, (S)-nicotine was effectively displaced by both isomers of nornicotine with IC50 values of approximately 0.9 nM (Zhang and Nordberg, 1993). An additional urinary metabolite in many species, including humans, is nicotine N-oxide. Although the N-oxide at large doses shared the pharmacological profile with nicotine in dogs (see Larson et al., 1961), this could be due to the parent nicotine regenerated from the N-oxide by gastrointestinal reduction as observed both in vitro and in vivo (Beckett et al., 1970; Crooks, 1993). The metabolic intermediate, the reactive alkylating agent ′ ′ nicotine- ∆1 (5 ) iminiumion (Nguyen et al., 1979), was recently implicated in the pharmacological and toxicological effects of nicotine in the brain (see Hukkanen et al., 2005). Interestingly, nicotine and (S)-anatabine are inhibitors of human hepatic CYP2A6 responsible for the conversion of nicotine to this iminium metabolite (Denton et al., 2004). The rapid metabolism of nicotine was illustrated in experiments with cats given 40   g/kg intravenous dose of [14C]nicotine (Turner, 1969). Tissues, including the brain, liver, kidneys, lungs, skeletal muscles, and stomach, showed maximum nicotine content in about 5 min after injection. Cotinine, appearing in the blood and liver within minutes of injection, was continuously transformed to other metabolites. Fifty-five percent of the radioactivity was excreted in the urine within 24 h, but only 1% of the radioactivity was unchanged nicotine. In 3 days, 70% of the injected radioactivity was excreted via urine, whereas feces contained less than 1% of radioactivity. Pharmacological actions and poisoning syndromes  The mammalian pharmacology of nicotine has been thoroughly reviewed (Benowitz, 1996; Brioni et al., 1997; Taylor, 2006) so it will only be discussed briefly.

130

Nicotine has a multitude of pharmacological and physiological effects, and the responses depend greatly on the dose and rate of absorption. In mammals, the alkaloid targets various receptor subtypes in both the central and the peripheral nervous systems (Maskos, 2007). In general, the initial stimulation is followed by depression. Both enantiomers of nicotine and nornicotine show the stimulant and depressant activities. For example, injection of nicotine into rats produces a biphasic effect on locomotor activity in a dose-dependent manner. First, ataxia is seen, which, at doses larger than 0.8 mg/kg, takes the form of prostration. About 10–20 min later, a stimulant effect can be observed. The stimulant effect was shown to involve nAChR-mediated dopamine release. The motor activation is followed by tremor, and seizures are also common. Typical stimulatory responses of the cardiovascular system to nicotine are increases in heart rate, myocardial contractility, and blood pressure. However, the drug can also slow the heart rate by paralyzing the sympathetic or by stimulating the parasympathetic cardiac ganglia. Nicotine also initiates the release of catecholamines in a number of isolated organs, causing additional cardial responses. The effect on the gastrointestinal tract is overall parasympathic stimulation, which results in increased bowel activity. Individuals not exposed previously to nicotine often experience nausea, vomiting, and diarrhea. The alkaloid has marked effects on the exocrine glands: the initial stimulation causing increased salivation is followed by inhibition of secretion. Neurochemical and behavioral assays suggest that nicotine metabolites are also pharmacologically active (see Crooks and Dwoskin, 1997). (d)  Toxicity to Laboratory Animals Acute toxicity  Over the past century, a plethora of data have accumulated on the toxicity of tobacco alkaloids. Table 3.2 lists representative acute toxicity data for various test species. Nicotine has a rapid contact and vapor action. With lethal exposures, renal failure, hypotension, paralysis, and coma may precede death, which is usually caused by respiratory failure due to both central paralysis and peripheral blockade of the muscles of respiration. The acute toxicity of cotinine is lower than that of the parent alkaloid, i.e., the intraperitoneal LD50 values for male and female mice are 2000 and 3000 mg/kg, respectively, while for nicotine hydrogen tartrate the respective LD50 values are 31 and 37 mg/kg (Riah et al., 1999). Nornicotine is generally less toxic than nicotine in most species but could be more toxic to some, depending on the mode of administration. In guinea pigs, for example, the subcutaneous LD50 values for nicotine and nornicotine are 32 and 28 mg/kg, respectively, but, for rats, the subcutaneous LD50 values are 23.5 mg/kg for both alkaloids (Negherbon, 1959; see also Larson et al., 1945). In mice, the intraperitoneal and intravenous LD50 values nornicotine are 21.7 and 3.4 mg/kg, respectively; in rabbits, the

Hayes’ Handbook of Pesticide Toxicology

respective LD50 values are greater than 13.7 and 3.0 mg/kg (Larson et al., 1945). The subcutaneous LD50 of nicotine N-oxide is 940 mg/kg in mice (Larson et al., 1961). Myosmine (Figure 3.2) is a trace component in Nicotiana species. In rats, the oral and intraperitoneal LD50 values are 1875 and 190 mg/kg, respectively (Ambrose and DeEds, 1946). Other pharmacological and biochemical effects Ruppert (1942) noted that a single dose of nicotine resulted in tolerance to a subsequent dose of nicotine (tachyphylaxis). In mice, for example, an intravenous sublethal dose of 0.8 mg/kg nicotine hemitartrate, when given 5 min before toxicity determination, raised the intravenous LD50 up to 20.8 mg/kg, which is 10 times higher than the corresponding value without pretreatment (Barrass et al., 1969; see also Benowitz et al., 1987). Nicotine N-oxide was almost as effective when given 40 min before nicotine. By activating elements of both the sympathetic and parasympathetic systems, acute administration of nicotine elicits a very wide spectrum of complex and sometimes unpredictable effects, including vasoconstriction, tachycardia, changes in blood pressure, increased tone and motor activity of the bowel, vomiting, etc. (see Taylor, 2006). Nicotine is both a substrate and an inhibitor of CYP enzymes involved in glucocorticoid and sex steroid biosynthesis. For example, nicotine, cotinine, and anabasine inhibit CYP-mediated adrenal aldosterone synthesis in rats (Skowronski and Feldman, 1994) and estrogen synthesis by human aromatase in vitro (Barbieri et al., 1986; see also Kadohama et al., 1993). Teratogenic, carcinogenic, and mutagenic effects  Nicotine and, to a lesser extent, cotinine were judged to be potential teratogens in a frog embryo teratogenesis assay (Dawson et al., 1988). Chronic oral administration of nicotine in drinking water (up to 100 g/ml) to gestating mice decreased the weight of the fetuses by up to 12% (Rowell and Clark, 1982). In rats, upon oral administration to pregnant rat dams nicotine (up to 17.5 g/ml) did not affect the body weight of pups but focal necrosis of the liver, presumably due to depressed superoxide dismutase activity and consequent oxidative stress, was noted at the highest nicotine dosage (Sheng et al., 2001). However, intraperitoneal injection to pregnant mice of nicotine sulfate at a dose of 1.67 mg/kg body weight during gestation days 6–15 reduced fetal crown–rump length and fetal head dimensions and 9.6% of the fetuses had clefts of the palate (Saad et al., 1990). Other studies have also suggested that interfering with the cholinergic signaling during fetal development leads to growth retardation and neural dysmorphology, as well as behavioral changes in the offspring (see, for example, Joschko et al., 1991). Recent investigations established that, in addition to polycyclic aromatic hydrocarbons, the carcinogenicity of

Chapter | 3  Pest Control Agents from Natural Products

131

Table 3.2  Acute Toxicity of Nicotine Animal

Route

LD50 (mg/kg)

Rat

oral

188a

Ambrose and DeEds (1946)

Rat

oral

50–60

Negherbon (1959)

Rat

oral

52.5

Lazutka et al. (1969)

Rat, female Rat Rat, female

oral oral dermal

83

b

90

b

285

Notes, other data

References

Gaines (1969) Bond et al. (1973) b

Gaines (1969)

c

Rat, male

sc

47

Rat, female

sc

37c

Holck et al. (1937)

Rat

sc

33.5

Negherbon (1959)

Rat

im

Rat, male

ip

14.6

Blum and Zacks (1958)

Rat

iv

2.8

Larson et al. (1949a)

Rat

iv

Mouse

oral

24

Heubner and Papierkowski (1938)

Mouse

oral

3.34

Lazutka et al. (1969)

Mouse

sc

16

Heubner and Papierkowski (1938)

Mouse

im

Mouse

iv

7.1

Larson et al. (1949b)

Mouse, female

iv

2.0e

Barrass et al. (1969)

Guinea pig

im

LD100  15 mg/kgd

Feurt et al. (1958)

Guinea pig

iv

LD100  4.5 mg/kgd

Negherbon (1959)

Rabbit

im

LD100  30 mg/kgd

Feurt et al. (1958)

Holck et al. (1937)

LD100  15 mg/kgd

LD100  1 mg/kgd

Feurt et al. (1958)

Negherbon (1959)

LD100  8.0 mg/kgd

Feurt et al. (1958)

Rabbit

dermal

50

Negherbon (1959)

Rabbit

ip

14

Larson et al. (1945)

Rabbit

iv

9.4

Larson et al. (1949b)

Cat

im

Cat

iv

Cat

iv

LD100  6.1 mg/kgd

Negherbon (1959)

Dog

im

LD100  15 mg/kgd

Feurt et al. (1958)

Dog

iv

Pig Goat Deer Cattle Horse Monkey

im im im im im im

LD100  9.0 mg/kgd 2.0

Feurt et al. (1958) Larson et al. (1949b)

5.0

Larson et al. (1949b) d

Feurt et al. (1958)

d

Feurt et al. (1958)

LD100  14 mg/kg LD100  13 mg/kg

d

Feurt et al. (1958)

d

Feurt et al. (1958)

d

Feurt et al. (1958)

d

Feurt et al. (1958)

LD100  9.0 mg/kg LD100  9.0 mg/kg LD100  8.8 mg/kg LD100  6.0 mg/kg

(Continued)

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132

Table 3.2  (Continued) Animal

Route

Pigeon

im

LD50 (mg/kg)

Notes, other data

References

LD100  9.0 mg/kgd

Feurt et al. (1958)

d

Negherbon (1959)

f

Lc50  316 mg/kg

Dawson et al. (1988)

Rainbow trout

LC50  4 ppm

Tomlin (2003)

Bluegill sunfish

LC50  5.45 ppm

U.S. EPA (2008a)

Daphnia pulex

LC50  0.24 ppm

Tomlin (2003)

EC50  72.9 ppmb,g

Seckar et al. (2008)

Frog Xenopus laevis embryo

Green alga, Selenastrum capricornutum

sc 96-h

96-h

LD100  40 mg/kg

a

Applied in aqueous solution adjusted to pH  7 with concentrated hydrochloric acid. Nicotine sulfate. Calculated from mortality data. d Approximate minimal lethal dose. e Nicotine hydrogen tartrate. f The LC50 value for cotinine was 4340 ppm in this assay. Both compounds caused various developmental malformations in the frog embryos; the LC50 values were 0.45 and 720 ppm for nicotine and cotinine, respectively. g Concentration inhibiting 50% of the algal growth rate. b c

tobacco and nicotine is associated with N-nitrosamines of secondary amines, such as N-nitrosonornicotine or certain nitrosaminoketones, either already present in tobacco or formed from nicotine during storage, processing, or biotransformation in the body (reviewed by Hecht, 2008). Cotinine was carcinogenic in rats (Truhaut et al., 1964). Neither nicotine sulfate (Moriya et al., 1983) nor nicotine base (Doolittle et al., 1995) was mutagenic in Salmonella typhimurium reversion-assay systems. In the latter study, cotinine and nicotine N-oxide were also nonmutagenic. Treatment  Artificial respiration alone before cessation of circulation or artificial respiration together with intracardial injection of epinephrine rescued nicotine-poisoned dogs (Franke and Thomas, 1936). Nicotine-induced convulsions can be blocked by certain anticholinergic drugs such as diethazin and diphenhydramine. The prostration response produced by injecting nicotine in rat brain was prevented by subcutaneous injection of mecamylamine (Shoaib and Stolerman, 1994). Tetraethylammonium chloride proved to be a useful antidote against acute doses of nicotine in mice (Andrews and Miskus, 1968). The behavioral, respiratory, and electrocortical effects of nicotine infused into the brain of fowl could be prevented or abolished by systemic or local administration of pempidine (Marley and Seller, 1974). Poisoning incidents  Application of a 5% aqueous nicotine solution on cattle to get rid of ectoparasites produced classical signs of nicotine toxicosis: accelerated respiration, salivation, sweating, diarrhea, and tremor. Of the 18 animals receiving nicotine, one died 2 h and another one

14 h after treatment. While 14 of the animals were symptomless on the second day, two showing serious symptoms were given digitalis, veratrine, and arecoline to recover (Kamarás, 1936). Several poisoning cases in dogs ingesting chewing tobacco or cigarette butts have been described (Hackendahl and Sereda, 2004; Sarkar, 2004; Vig, 1990). Treatments included conjunctival administration of apomorphine to induce vomiting followed by activated charcoal orally or intramuscular administrations of the respiratory stimulant nikethamide and atropine, to reverse cholinergic excess. Lethal poisoning of mules that were feeding hay contaminated with drippings of tobacco stalks and leaves previously hung over the hay has recently been described (Sanecki et al., 1994). Symptomatic and antidotal treatment, including atropine in combination with the antispasmodic drug memantine, was ineffective at the advanced paralytic phase of the toxicosis. (e)  Toxicity to Humans Effect on reproduction  Maternal smoking during pregnancy has adverse consequences not only for the mother but her fetus and the newborn (reviewed by Lambers and Clark, 1996). Nicotine increases spontaneous abortions or premature delivery rates and decreases birth weight. Nicotine was shown to concentrate in fetal blood, amniotic fluid, and breast milk, causing various effects discussed earlier in both the fetus and the neonate. Mostly due to its cholinergic properties, the alkaloid affects embryonic neural development through interfering with neural cell replication and differentiation, disrupting axonogenesis and synaptogenesis, evoking or preventing apoptosis, and

Chapter | 3  Pest Control Agents from Natural Products

ultimately compromising synaptic function in multiple brain regions (Pauly and Slotkin, 2008). Poisoning incidents  Accidental poisonings by solutions and vapors of nicotine were rather common in the early part of the 20th century when this insecticide was extensively used (Faulkner, 1933; Wehrlin, 1938) and fatal or serious reactions have been reported for various types of exposures, including dermal absorption. Symptoms of acute nicotine poisoning occur rapidly and include nausea, salivation, abdominal pain, vomiting, diarrhea, cold sweat, headache, dizziness, disturbed hearing and vision, mental confusion, and overall weakness. The blood pressure falls and faintness and prostration occur. Breathing becomes difficult; the pulse is weak, rapid, and irregular. Collapse may be followed by convulsions. Death may result within a few minutes from respiratory failure (Taylor, 2006). In dermal poisoning, however, the onset of symptoms could be delayed by several hours (Benowitz et al., 1987). The often-cited acutely fatal dose of nicotine for an adult is about 60 mg (one drop) of nicotine, yet individuals have ingested larger quantities and recovered (see Franke and Thomas, 1936). Green-tobacco sickness is an occupational illness occurring regularly in tobacco fields and is caused by the dermal absorption of nicotine while harvesting (Arcury et al., 2008; McBride et al., 1998). Symptoms usually occur several hours after harvesting of wet green tobacco begins and last no longer than 24 h if contact with tobacco leaves is avoided. The initial headache and nausea usually lead to severe vomiting, pallor, and prostration. Smoking appears to protect against the illness. Contact dermatitis is another occupational disease of tobacco farmworkers (see, for example, Abraham et al., 2007). The distribution of dermatitis (flexor and surfaces of the arms, axilla, and torso) is consistent with the practice of holding the tobacco leaves under the arm while harvesting but the involvement of plant growth regulators ethephon or maleic hydrazide used in tobacco cannot be ruled out. Rogers et al. (2004) describe a recent nicotine insecticide poisoning that involved an adolescent who ingested an unknown amount of “Black Leaf 40” solution and was presented with generalized seizures, a rectal temperature of 33.8°C, a heart rate of 157 beats/min, blood pressure 170/108 mmHg, and a spontaneous respiration rate of 6 breaths/min. In spite of aggressive intensive care support, the patient showed minimal improvement and even 6 months later showed signs of encephalopathy and was dependent on gastric tube feedings. Weiss (1996) described a case of acute nicotine toxicosis for a patient hospitalized for lithium therapy for bipolar disorder. The patient wanted to give up smoking and was given a transdermal nicotine patch (21 mg/day). He still smoked intermittently and 4 days later complained of nausea, diaphoresis, and hand tremor. The symptoms were

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first misdiagnosed as lithium toxicity but soon corrected as a typical case of nicotine toxicity and the transdermal nicotine treatment discontinued. It may also be noted that several traditional medicines contain tobacco preparations that could cause nicotinepoisoning (see, for example, Davies et al., 2001; Oyebola, 1983). Garcia-Estrada and Fischman (1977) described an unusual case of nicotine poisoning by homemade tobacco enema for which an aqueous concoction of 5–10 cigarettes was used. The patient’s hypotension and bradycardia were reversed about 10 h after the enema by intravenous atropine. The universal availability of nicotine has made it a common means of suicidal poisoning as illustrated by several cases in the literature. In a fatal suicide case, described by Lavoie and Harris (1991), a 17-year-old smoker ingested an estimated 5 g of a concentrated nicotine insecticide solution then vomited and collapsed, pulseless. Following immediate CPR, the initial asystole was converted to ventricular fibrillation with intravenous epinephrine. Subsequent orotracheal intubation and defibrillation produced a sinus tachycardia and the patient was placed on lidocaine for premature ventricular contractions. In the emergency room, mechanical ventilation and dopamine infusion were applied. Gastric lavage with normal saline followed by charcoal was implemented. The severe convulsions that occurred were controlled with diazepam, dilantin, mannitol, and dexamethason. The urine nicotine screen was positive and the serum level was 13,600 ng/ml, enough to be lethal [see also Moriya and Hashimoto, 2005; the average peak level for smokers is 49 ng/ml (Russell et al., 1976)]. Computer tomography showed cerebral edema and electroencephalogram (EEG) revealed no cortical function. On the second day of hospitalization, intractable hypotension set in and the patient died 64 h after ingestion. Besides insecticide formulations, the traditional sources of nicotine were snuff or cigarette butts (Saxena and Scheman, 1985). The recently developed and easily accessible transdermal nicotine patches, which typically contain 7–114 mg of the alkaloid, are novel sources of the poison for suicide attempts (Engel and Parmentier, 1993; Kemp et al., 1997) or for accidental intoxication among children (Woolf et al., 1997). Treatment  Early removal of nicotine, aggressive respiratory support, and treatment of shock are important countermeasures (Franke and Thomas, 1936; Taylor, 2006). In the absence of seizures, vomiting should be induced; gastric lavage could also be performed. Activated charcoal is a valuable adjunct in neutralizing ingested nicotine in the stomach, but tannic acid solution, although useful in precipitating some other alkaloids (e.g. strychnine), is of little use for this alkaloid (Hayes, 1975). Potassium permanganate diluted 10,000-fold in the lavage fluid can also be used. Alkaline solutions, which facilitate absorption,

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should be avoided. Patients exhibiting convulsions may require sedation (e.g. by intravenous diazepam) and vasopressor drugs if hypotension fails to respond to the usual therapy. Patients surviving for more than a few hours after ingesting nicotine are likely to recover because nicotine is detoxified fairly rapidly. Pathology  In contrast to acute poisoning cases where there are no specific pathological changes attributable to nicotine, chronic exposure to nicotine is implicated in the pathogenesis of coronary and peripheral vascular disease, chronic lung disease, cancer, and various endocrine disturbances (Benowitz, 1986, 1996).

3.2.1.3  Anabasine (a)  Introduction Anabasine (Figure 3.2), a close structural relative of nicotine, was isolated from the toxic Asian plant Anabasis aphylla (Chenopodiaceae) by Orechoff and Menschikoff in 1931. The same year, Smith isolated this compound as an insecticidal trace constituent of synthetic dipyridyl oil and named it “neonicotine.” The anabasine content of A. aphylla could be as high as 2.6% and the alkaloid is also present in Nicotiana species (e.g. the wild tree tobacco, N. glauca). Recently, anabasine has also been isolated from certain worms and ants (see Leclercq et al., 2001). Anabasine sulfate was a widely used botanical insecticide in the former Soviet Union until the 1970s. (b)  Identity, Physicochemical Properties, and Uses IUPAC name: (S)-3-(piperidin-2-yl)pyridine. Chemical Abstract name: 3-[(2S)-(2-piperidinyl)pyridine. CAS Registry Number: (S)-anabasine: [494-52-0]; racemic anabasine: [13078-04-1]. Empirical formula: C10H14N2; molecular weight: 162.2. Anabasine is a clear liquid with a boiling point of 277– 280°C. Its freezing point is 9°C. The density of anabasine is 1.048 g/cm3 at 20°C. The alkaloid is miscible with water and most organic solvents. Anabasine from A. aphylla is 20 levorotatory [ ]D  82.20, but the alkaloid obtained for practical purposes from the American N. glauca is mostly racemic. Stability  Anabasine is somewhat more stable than nicotine, but in air it also turns brown. (c)  Biological Properties Mode of action  Similar to nicotine, anabasine is also a nAChR agonist alkaloid. (d)  Toxicity to Animals Acute toxicity  Anabasine was more toxic than nicotine when given intravenously to rabbits: the minimal fatal doses

Hayes’ Handbook of Pesticide Toxicology

for anabasine and nicotine were 3 and 9 mg/kg, respectively (Haag, 1933). For guinea pigs, the difference between anabasine and nicotine was less with respective fatal doses of 22 and 26 mg/kg on subcutaneous administration. In the mouse, the intravenous LD50 of (S)-anabasine was 16 mg/kg thus being less toxic than nicotine (Lee et al., 2006). Acute anabasine poisoning symptoms are the same as those for nicotine: initially increased respiration, then hyperexcitability, muscular twitching, followed by depression of respiration and complete muscular paralysis. Anabasine, however, was found to be less excitatory and more depressing than nicotine. The animals completely recovered from sublethal doses of anabasine in about one hour. The teratogenicity of N. glauca plants and its primary alkaloid, anabasine, in pigs, sheep, goats, calves, or chickens was extensively studied (Bush and Crowe, 1989). The principal defects in piglets were deformities of fore and rear limbs (arthrogryposis) and cleft palate that could be induced at 2.6 mg/kg doses of anabasine twice daily during the 43rd to 53rd days of gestation (Keeler et al., 1984). In goats, a similar study indicated that anabasine dramatically decreased fetal movement, as evidenced by ultrasound examination, with concomitant palatal clefting (Weinzweig et al., 2008). Poisoning incidents  Habib et al. (1974) reported that an anabasine-containing plant, Haloxylon persicum (Chenopodiaceae), growing wild in Saudi Arabia, caused death among grazing animals. A case of N. glauca toxicosis in a herd of cattle ingesting leaves of the tree tobacco was reported by Plumlee et al. (1993). Symptoms of poisoning were ataxia, depression, anorexia, and mild colic. Necropsy of one of the dead animals revealed edematous and congested lungs with peracute aspirative pneumonia. Histology indicated a mild, multifocal, suppurative rumenitis. Anabasine was found in the liver (2 ppm), urine (3 ppm), and rumen contents (20 ppm). Human poisoning incidents  A poisoning case due to homemade enema with an aqueous solution of anabasine sulfate was reported by Danilin and Shabaeva (1968). The initial symptoms 30 min after administration of the enema were general weakness, nausea, cardiac disturbances with an arterial blood pressure of 80/50, pulse rate of 42 per min, a body temperature of 35.2°C, and acrocyanosis. Gastric lavage was performed. After an additional 35 min, respiration became severely depressed, and the patient sweated and lost consciousness. Resuscitation and intubation were employed and subcutaneous atropine and nikethamide were given and the patient’s condition slowly improved over the next 6 h. Anabasine was subsequently identified in the urine but not in the gastric lavage fluid because the alkaloid had been removed by defecation that immediately followed the enema. Complete recovery took several days.

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A fatal anabasine poisoning due to ingestion of N. glauca leaves was described by Castorena et al. (1987). A young adult male was found in the field where the plant was also identified. Autopsy revealed only moderate congestion and edema of both lungs. Anabasine concentration of the gastric content, urine, and blood was 113.4, 73.8, and 1.15 mg/l, respectively. The heart, kidney, brain, and lung contained 10.4–15.8 mg/l of the alkaloid. An extract of the leaves collected in the field contained 2.0 mg/g of anabasine. Similar poisoning cases have been described by Steenkamp et al. (2002). Fresh N. glauca leaves were mistaken for the edible Amaranthus hybridus, used as “spinach” (marog) in a traditional meal in southern Africa. The ingestion of the meal led to two fatalities. The anabasine concentrations in these cases were relatively low (9.3 and 19.1 mg/l in the stomach) but both deceased vomited extensively before dying from neuromuscular blockade resulting in respiratory depression. While several other nonfatal poisoning cases due to ingestion of N. glauca leaves are known (Mellick et al., 1999), a recent case described by Murphy et al. (2006) concerns a 3-month-old child dermally exposed to anabasine from N. glauca leaf placed on the baby’s abdomen by her mother for several hours as a topical folk remedy to treat constipation. Upon presentation to the hospital, main symptoms were apnea, absent gag reflex, miosis, flaccid paralysis, and peak blood pressure of 139/101 mmHg. After intubation and symptomatic treatment (e.g. intravenous labetalol to reduce hypertension), the child’s vital signs improved only 12 h after presentation to the hospital and head computer tomography revealed mild brain atrophy. Anabasine concentration in the urine was 30,690 ng/ ml at this time. The child became symptomless 36 h after presentation when extubation was performed. Treatment  Treatment is the same as that for nicotine poisoning.

3.2.1.4  Rotenone (a)  Introduction Roots of the East Asian Derris (tuba) plants, particularly of Derris elliptica and D. chinensis (Leguminosae), have long been used to stupefy fish for easy collection and also as insecticides and hunting aids. In South America, Lonchocarpus utilis, L. urucu and L. nicou (cubé, timbo, and barbasco), as well as preparations from Tephrosia, Mundela, and Millettia species, have been used for the same purpose. The major and most studied bioactive principle of the tropical plants is rotenone, first isolated from L. nicou by Geoffroy in 1895. The rotenone content in D. elliptica and L. utilis roots ranges from 3 to 11%. The structure of rotenone (Figure 3.3) was established by several research groups simultaneously in 1932 (reviewed by Crombie, 1963; LaForge et al., 1933). Additional notable

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O O

O

3

H 12a O

6a H

O

O H

rotenone

6' 8' 7'

Figure 3.3  Structure of rotenone.

rotenoids are deguelin, ellipton, malaccol, sumatrol, tephrosin, and -toxicarol (Crombie, 1963; Fukami and Nakajima, 1971) as well as the recently identified oxahomologs (Fang and Casida, 1997, 1999). (b)  Identity, Physicochemical Properties, and Uses IUPAC name: (2R,6aS,12aS)-1,2,6,6a,12,12a-hexahydro2-isopropenyl-8,9-dimethoxychromeno[3,4-b]furo[2,3h]chromen-6-one. Chemical Abstract name: (2R,6aS,12aS)-1,2,12,12a-tetrahydro-8,9-dimethoxy-2-(1-methylethenyl)-[1]benzopyrano[3,4-b]furo[2,3-h][1]benzopyran-6(6aH)-one. CAS Registry Number: [83-79-4]. Empirical formula: C23H22O6; molecular weight: 394.4. Physicochemical properties  The melting point of rotenote in its orthorhombic form is 165–166°C; in its dimorphic form, the melting point is 185–186°C. At 20°C, rotenone is practically insoluble in water (0.2 mg/l) (see, however, Loeb and Engstrom-Heg, 1970) but readily soluble in chloroform (472 g/l) and acetone (66 g/l) and only slightly soluble in ethyl ether (4 g/l) and ethyl alcohol (2 g/l). Rotenone is highly lipophilic with a log P of 4.26 (Gingerich and Rach, 1985). It is levorotatory: [ ]20 D  228 (c  2.22 in benzene). Stability  Derris and cubé preparations as well as rotenone are stable on storage. When exposed to light and air in the field, however, rotenone decomposes rapidly, losing toxicity within days. Upon exposure to light and air, rotenone solutions become successively yellow, orange, and deep red (Cheng et al., 1972). Rotenone is readily isomerized by bases, even by alkaline glass surfaces. After use in fish management programs, residues of the easily oxidizable rotenone can be neutralized by dilute solutions of potassium permanganate (Lawrence, 1956) or chlorine. The decomposition of rotenone in solution, on plants and glass surfaces was studied in detail by Cheng et al. (1972), who found that irradiation of rotenone in

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oxygenated methanol with ultraviolet (UV) light yielded a mixture of over 20 photodegradation products with lower mammalian toxicity than the intact parent compound. The most toxic and prevailing degradate isolated was the 12ahydroxylated derivative 6a,12a-rotenolone, having an intraperitoneal LD50 of 4.1 mg/kg to male mice (see also Fukami et al., 1967). The corresponding LD50 for rotenone was 2.8 mg/kg in this study. Environmental conditions affect the rate at which rotenone is degraded. Cabizza et al. (2004) found that 1–20% of the rotenone content of various cubé resin formulations was lost after 6 months of normal storage (darkness and room temperature). For other rotenoids, including deguelin and tephrosin, larger changes were observed. In photodegradation experiments, the half-lives of rotenone in these formulations were 18–27 min. Analysis of honey collected from beehives treated weekly for 1 month with a liquid rotenone formulation containing 3% rotenone to protect bee colonies from Varroa mite infestation showed average rotenone concentration of 0.11 mg/kg (Jiménez et al., 2000). Studying the degradation of rotenone in different soils, Cavoski et al. (2008) reported half-lives of 21–25 and 5–8 days at 10 and 20°C, respectively. In ponds and water reservoirs, the half-life of rotenone is 15–40 and approximately 84 h at 20–24 and 0–15°C, respectively (Finlayson et al., 2000). The toxicity of a 2-ppm aqueous rotenone solution is lost in 3 days at 20°C but in 11–16 days at 11°C, as determined by the survival time of the roach, Rutilus rutilus (Meadows, 1973). Formulations and uses  Cubé, derris, and tuba preparations are commercialized alone or in combination with other botanical insecticides and PB synergist. Products containing pure rotenone as the sole active ingredient are used as agricultural insecticides and acaricides in orchards and vegetable cultivations as well as to control fire ants and household insects. Rotenone is also used to treat scabies and head lice on humans as well as various ectoparasites on livestock and pet animals. It is formulated as dust, wettable powder, and emulsifiable or soluble concentrates typically containing 0.4–8% of the active ingredient. In fishery management, powder, liquid or poisoned bait formulations are used to kill unwanted fish in ponds normally at 0.02–0.25 ppm calculated final rotenone concentrations (Finlayson et al., 2000; Rayner and Creese, 2006). Rotenone has also been used for small-scale sampling of fish to assess biodiversity and biomass. (c)  Biological Properties Mode of insecticidal action  Rotenone is a prototype inhibitor of the respiratory chain of mitochondria and acts by inhibiting electron transport at reduced nicotinamide adenine dinucleotide (NADH):ubiquinone oxidoreductase (Complex I). Complex I of the well-characterized

Hayes’ Handbook of Pesticide Toxicology

bovine heart mitochondrial membranes contains one flavin coenzyme and at least four iron-sulfur centers and is linked to additional polypeptide electron carrier complexes that involve iron-sulfur proteins and flavin or cytochrome coenzymes (Pilkington et al., 1993). The inhibition of Complex I ultimately results in loss of oxidative phosphorylation, so that adenosine 5-triphosphate (ATP) levels fall rapidly and cell death ensues. An additional effect of the inhibition of Complex I is the increase in the levels of reactive oxygen species (ROS) and nitrogen oxide (NO), free radical production and consequential lipid peroxidation in mitochondrial and associated cellular components leading to apoptosis (programmed cell death) (see, for example, Li et al., 2003). A recent comparative study by Salehzadeh et al. (2002) showed that rotenone effectively inhibited multiplication of several insect and mammalian cells in vitro at submicromolar concentrations with some selectivity against insect cell lines. Azadirachtin, a structurally unrelated botanical insecticide (see later), was more potent and selective with EC50 values being in the submicromolar concentrations for Sf9 ovary cells from Spodoptera frugiperda and the C6/36 cell line from Aedes albopictus larvae. Interestingly, pyrethrum extract stimulated all cell culture growth at submicromolar concentrations while nicotine was essentially ineffective below 106 M, the highest concentration tested. Complex I is also the target of other structurally different natural and synthetic pesticides (Hollingworth, 2001; Sparks and DeAmicis, 2007; Walter, 2007). Papaverine, a noninsecticidal spasmolytic, was also a respiratory inhibitor in vitro with the same type of activity as rotenone (Fukami, 1976; Santi et al., 1963, 1964). Rotenone has feeding deterrent activity against stored product insects (Nawrot et al., 1989; see also Boeke et al., 2004a). Metabolism and excretion  The selectivity of rotenone originates mainly from differences in the detoxification rates in various organisms and not from differences in the target enzyme system. Rotenone is effectively transformed by microsomal mixed-function oxidases in the liver, and the metabolic pathways as well as the chemical and biological nature of the products are well defined for mammals, insects, and certain fish (Fukami et al., 1969; Yamamoto et al., 1971). The principal metabolites formed both in vivo and in vitro from rotenone for these species are basically the same and are, in decreasing order of mammalian toxicity, as follows: 8-hydroxyrotenone, 6a,12a-rotenolone, 6a,12arotenolone, and 6,7-dihydro-6,7-dihydroxyrotenone (Fukami et al., 1967). The hydroxylated metabolites also have reduced inhibitory activity to insect and rat liver mitochondria (Horgan et al., 1968). The formation of watersoluble conjugates, which were more abundant in mammalian than in insect tissues, was also noted in these studies. A phenolic metabolite resulting from 3-O-demethylation

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was also identified (Unai et al., 1973). Synergists that block oxidative metabolism enhance the toxicity of rotenone to both insects and vertebrates. Rotenone was shown to be susceptible to microbial oxidation (see Sariaslani and Rosazza, 1985). Early experiments with rabbits and dogs fed with rotenone indicated retention and/or metabolism of the toxicant (Ambrose and Haag, 1937). No intact material was obtained from the urine, but the feces contained rotenone for at least 8 days after administration. A somewhat different picture emerged for mice (Fukami et al., 1967): 48 h after treatment with 14C-labeled rotenone, 20% of the radioactivity was in the urine, 0.3% was expired, 5% remained in the body, and the rest was in the feces. When the fate and distribution of 14C-labeled rotenone in different organs were followed in mice (Fukami et al., 1967), 21.6% of the radioactivity was found in the small intestine, 19.5% in the urine, and 4.4% in the liver. In the urine, more than 82% of the product was water soluble, 17% was 6,7-dihydro-6,7-dihydroxyrotenone, and only 1% was unchanged rotenone. In the liver and small intestine, 6,7-dihydro-6,7-dihydroxyrotenone was the predominant metabolite. However, water-soluble products also formed in almost equal amount, and 10–16% rotenone and a few percentages of other monohydroxylated metabolites were also present. In a similar experiment with carp, about 20% of the administered rotenone could be recovered from the analyzed tissues, and the bulk of the metabolites (45%) consisted of unidentified water-soluble products. The metabolism of rotenone in fish was studied in detail. For example, the half-life of rotenone in the head, viscera, and carcass of bluegills was about 22, 11, and 28 days, respectively, and the major metabolites identified were rotenolone and 6,7-dihydro-6,7-dihydroxyrotenone (Gingerich and Rach, 1985; see also Gingerich, 1986; Rach and Gingerich, 1986). Notable differences exist among fish species both in vivo and in vitro: Carp, rainbow trout, and bluegill produced the hydroxylated metabolites in different relative ratios (Erickson et al., 1992; Gingerich and Rach, 1985). The fate of intravenous [14C] rotenone (120 g/kg) in rainbow trout was studied by Gingerich (1986). After 20 min of injection, over 98% of the radioactivity was cleared from the plasma and started to accumulate in the heart ventricle, lateral-line red muscle, and posterior kidney, tissues that are highly dependent on aerobic metabolism, and in the liver, pyloric caeca, and small intestine. After 18 h, over 40% of the radioactivity of the liver, kidney, and muscle tissues was associated with the mitochondrial fractions. The estimated half-life of rotenone in the whole body was 68.5 h, and the major metabolite was 6,7-dihydro-6,7-dihydroxyrotenone. Biochemical effects and pharmacology  There are numerous reports on the physiological effects of rotenone on

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contractile responses in isolated guinea pig muscle preparations, including negative inotropic and chronotropic effects on atria, blocking of barium chloride- and bradykinininduced ileum contraction, inhibition of epinephrine-induced contraction of seminal vesicles, as well as antagonism of the chemically induced slow-reacting substance of anaphylaxis in isolated ileum (see, for example, Ashack et al., 1980; Haley, 1978, and references therein). Because the dysfunction of Complex I has been implicated in the pathogenesis of Parkinson’s disease, binding of rotenone and its interference at this site with the prototypical parkinsonism-inducing agent 1-methyl-4phenylpyridinium (MPP) were examined in several laboratories. Heikkila et al. (1985) reported that systemic injection of rotenone into rat brain resulted in nigrostriatal cell death similar to that observed for MPP. Later, Ramsay et al. (1991) showed that rotenone and MPP share a common binding site at Complex I supporting the validity of the mitochondrial inhibition hypothesis for chemically induced parkinsonism. It has also been demonstrated that the molecular target of rotenone and MPP differ from that of paraquat, which does not inhibit Complex I though it causes neurodegeneration of dopaminergic neurons (Richardson et al., 2005). Chronic infusion of 2–3 mg/kg daily doses of rotenone to rats caused parkinsonian behavior (bradykinesia, rigidity, postural instability and tremor), as well as specific nigrostriatal dopaminergic cell degeneration as indicated by reduced immunohistochemical staining of tyrosine hydroxylase (TH), dopamine transporter, and vesicular monoamine transporter type 2 (VMAT2) proteins (Betarbet et al., 2000). Conversely, upon histologic evaluation of the brains of rats given daily intravenous infusions of 10–18 mg/kg rotenone for 7–9 days, Ferrante et al. (1997) found selective bilateral lesions within the striatum and globus pallidus but not in the substantia nigra, suggesting that factors other than Complex I, for example, selective dopamine/MPP transporters, were also involved in the induced neurotoxicity. Indeed, rotenone was shown to selectively and efficiently block the uptake of dopamine in nigrostriatal dopaminergic neurons (Marey-Semper et al., 1993). Furthermore, Watabe and Nakaki (2008) have recently shown that at the relatively high 0.4 M concentration rotenone inhibits human VMAT2, responsible not only for packing dopamine and other monoamines into monoamine-containing neurons and into synaptic vesicles at nerve terminals but also for providing neuroprotection against neurotoxins. Inhibition of VMAT2 leads to dopamine redistribution from vesicles into cytosol where, upon oxidative activation, the accumulated neurotransmitters act as a neurotoxin causing neurodegeneration. It was also shown that the elevated NO production induced by rotenone results in nitration of tyrosine residues of both VMAT2 and TH. The NO synthase inhibitor N-nitro-l-arginine methyl ester and the ROS scavenger N-acetylcysteine

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attenuated rotenone-induced dopamine redistribution as well as apoptosis of dopaminergic neurons. Results of subsequent studies (see, for example, Caboni et al., 2004; Lapointe et al., 2004; Schmidt and Alam, 2006; Sherer et al., 2003; reviewed by Bové et al., 2005; Höglinger et al., 2006) corroborate that repeated systemic application of relatively high doses of rotenone causes selective oxidative damage to nigrostriatal dopaminergic regions both in vitro and in vivo and also suggest that this botanical insecticide/piscicide is an atypical rather than idiopathic rodent model of parkinsonism. However, the available animal and epidemiological data do not provide convincing evidence to support a casual association between rotenone exposure and Parkinson’s disease in humans (Brown et al., 2006). Sherer et al. (2007) have recently reported that besides rotenone other insecticides/acaricides acting at Complex I (fenpyroximate, fenazaquin, tebufenpyrad, and especially pyridaben) can cause oxidative damage and cell death in vitro. The antioxidants -tocopherol and coenzyme Q10 were protective against the neurotoxicity. The oxidative stress initiated by the Complex I inhibitory effect of rotenone in various cell types could be prevented by l-deprenyl (Saravanan et al., 2006), l-DOPA (Schmidt and Alam, 2006), lamotrigine (Kim et al., 2007), melatonin (Saravanan et al., 2007), vitamin E (Zhang et al., 2001), and H2S (Hu et al., 2009). Simultaneous administration of the antioxidant nootropic agent meclofenoxate (centrophenoxine) attenuated behavioral dysfunction associated with the dopaminergic toxicity of chronic rotenone treatment in rats (Nehru et al., 2008). Furthermore, local intraocular administration of the NMDA receptor antagonist memantine protected rotenone-induced retinal degeneration in mice indicating a critical interaction between excitatory cell membranes and mitochondria (Rojas et al., 2008). The fatty acid changes seen in the liver after chronic rotenone feeding (Haag, 1931) were attributed to the accumulation of long-chain acyl-coenzyme A (acyl-CoA), for example palmitoyl-CoA as shown for rabbit heart mitochondria (Hull and Whereat, 1967). In cultured rat hepatocytes, the mitochondrial membrane permeability transition caused by rotenone via this mechanism could be prevented by l-carnitine presumably converting palmitoyl-CoA to palmitoyl carnitine, which does not induce the transition (Pastorino et al., 1993). Furthermore, rotenone-induced anoxic death in rat hepatocytes could be prevented in vitro by either l-carnitine or the immunosuppressive drug cyclosporin A (Pastorino et al., 1993), as well as by removing Ca2 from the culture medium (Pastorino et al., 1995). According to Terzi et al. (2004), the mucolytic antioxidant erdosteine prevented rotenone-induced liver injury in rats presumably by inhibiting lipid peroxidation and xanthine oxidase activity. In the liver of rats treated with rotenone  erdosteine, increased catalase and superoxide

Hayes’ Handbook of Pesticide Toxicology

dismutase activity was observed but the NO level was unchanged compared to rotenone-only controls. A recent interesting work by Varma et al. (2007) showed that rotenone suppressed neuronal death in a rat striatal model of human Huntington’s disease (HD) in vitro and also in animal models of HD in vivo. The mechanism by which rotenone and some other mitochondrial inhibitors, including 2,4-dinitrophenol, rescue cell death involves caspase inhibition and activation of prosurvival signaling. Rotenone also activated extracellular signal-regulated kinase (ERK) and Akt protein kinase, both known to enhance neuronal cell death survival in HD models. Cytotoxicity and mutagenicity  The cytotoxic and antitumor activity of rotenone is well established, but its carcinogenic potential has been a matter of controversy. On the one hand, Gosálvez (1983) found that chronic administration of rotenone, given either daily intraperitoneal doses of 1.7 mg/kg for 42 days or daily oral doses of 13.5 mg/kg for 45 days, induced mammary gland tumors in female albino or Wistar rats. In follow-up studies, daily intraperitoneal doses of rotenone at 7.1–9.1 mg/kg for 42 days also induced tumors in female Wistar rats. The authors, however, suggested that the observed effect might be associated with diets deficient in minerals and vitamins, especially riboflavin. On the other hand, several studies showed lack of carcinogenicity for rotenone (California EPA, 1997); in particular, a detailed investigation by Greenman et al. (1993) could not reproduce previous experimental results on mammary gland carcinogenicity. Microscopic investigations with cultured mammalian cells showed that rotenone delayed cell progression in all phases of cell development and reversibly inhibited mitotic spindle microtubule assembly (Barham and Brinkley, 1976; Marshall and Himes, 1978; see also Ren and Feng, 2007). Rotenone was found to be cytotoxic to a number of human cancer cell lines, including solid tumor types, with ED50 values ranging from 0.05 to 0.15 g/ml (Blaskó et al., 1989; see also Konoshima et al., 1993; Salehzadeh et al., 2002). Furthermore, on short-term feeding, cubé (Hansen et al., 1965) and rotenone (see Wang et al., 2004) reduced either spontaneous or chemically induced tumors in rodents. Dietary rotenone also inhibited chemically induced tongue carcinogenesis in rats (Tanaka et al., 2002). Fang and Casida (1998) recently showed that rotenone and 28 other rotenoids blocked phorbol-ester-induced ornithine decarboxylase activity in human epithelial breast cancer cells in vitro, which correlated well with the inhibition of bovine heart NADH:ubiquinone oxidoreductase activity by these compounds. Rotenone treatment did not increase chromosome aberrations in Chinese hamster ovary cells (California EPA, 1997). In another test with human lymphocyte cultures, rotenone increased the frequency of binucleated micronucleated cells and caused a delay in cell cycle but did not

Chapter | 3  Pest Control Agents from Natural Products

influence chromosome aberrations and sister–chromatid exchanges (Guadaño et al., 1998). Rotenone was nonmutagenic in bacterial reversion tests (California EPA, 1997; Moriya et al., 1983). (d)  Toxicity to Animals Acute toxicity  Results of various toxicity studies with derris and cubé preparations were published (Ambrose and

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Haag, 1936, 1937, 1938; Cutkomp, 1943; Mathews and Lightbody, 1936; Negherbon, 1959), as were data for individual rotenoids (Fukami and Nakajima, 1971; Negherbon, 1959). Representative acute toxicity data of purified rotenone are shown in Table 3.3 (see also Ling, 2003). Rotenone was reported to be emetic to dogs (California EPA, 1997; Lightbody and Mathews, 1936) and lethal to pigs at 3.7 mg/kg oral dose (see Oliver and Roe, 1957).

Table 3.3  Acute Toxicity of Rotenone Animal Rat Rat Rat

Assay oral oral

LD50 (mg/kg) a

25

a

150

Notes, other data

References

In olive oil

Lightbody and Mathews (1936)

As fine crystals

Lightbody and Mathews (1936) a

LD70  600 mg/kg

Ambrose and Haag (1937)

b

oral

Rat, male

oral

60

In acetone

Santi and Tóth (1965)

Rat, male

oral

102

In corn oil

U.S. EPA (2007)

Rat, female

oral

39.5

In corn oil

U.S. EPA (2007)

Rat, male Rat

b

ip

1.6

iv

a

Santi and Tóth (1965)

6

Lightbody and Mathews (1936) b

Rat, male

iv

0.2

Santi and Tóth (1965)

Mouse

oral

350

Tomlin (2003)

Mouse, male Guinea pig Guinea pig Guinea pig Guinea pig Guinea pig Rabbit Rabbit Rabbit Rabbit

ip oral oral

c

2.8

Cheng et al. (1972)

a

In olive oil

a

In starch paste

12

50

Lightbody and Mathews (1936) Lightbody and Mathews (1936) a

oral

Ambrose and Haag (1937)

LD70  60 mg/kg

d

sc

Haag (1931)

MLD  16 mg/kg

ip

MLD  2.0 mg/kg

oral

d

Haag (1931) d

Haag (1931)

a

Ambrose and Haag (1937)

MLD  1500 mg/kg

oral

LD70  3000 mg/kg d

sc

Haag (1931)

MLD  20 mg/kg

iv

MLD  0.35 mg/kg

d

Haag (1931)

d

Haag (1931) Haag (1931)

Cat

iv

MLD  0.65 mg/kg

Dog

iv

MLD  0.65 mg/kgd

Chicken

oral

996c

Cutkomp (1943)

Pheasant

oral

c

850

Cutkomp (1943)

Pigeon

oral

100c

Cutkomp (1943)

Pigeon

iv

Eastern robin

oral

195c

Cutkomp (1943)

Mallard duck

oral

2500

Ling (2003)

Frog Rainbow trout

48-h acute

MLD  1.0 mg/kgd

Haag (1931)

LC50  2 ppm

Haag (1931)

LC50  28 ppb

Bridges and Cope (1965) (Continued)

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Table 3.3  (Continued) Animal

Assay

Rainbow trout

24-h acute

LD50 (mg/kg)

Notes, other data

References Ling (2003)

LC50  2.2 ppb e

Rainbow trout

48-h acute

LC50  1.2–5.8 ppm, various derris formulations

Tooby et al. (1975)

Rainbow trout

48-h acute

LC50  0.34–0.47 ppm,f various derris formulations

Tooby et al. (1975)

Rainbow trout

96-h acute

LC50  5.8 ppb

Cheng and Farrell (2007) g

White perch

24-h acute

LC100  150 ppb

Wujtewicz et al. (1997)

Chinook salmon

24-h acute

LC50  5.6 ppb

Ling (2003)

Common carp

24-h acute

LC50  4.2 ppb

Ling (2003)

g,h

Crayfish

24-h acute

LC0  3.0 ppm

Daphnia pulex

1-h acute

LC100  250 ppb i

Wujtewicz et al. (1997) Negherbon (1959)

Northern snakefish

1-h acute

LC100  75 ppb

Lazur et al. (2006)

Bluegill

96-h acute

LC50  23 ppb

Bridges and Cope (1965)

Bluegill

24-h acute

LC50  26 ppb

Gingerich and Rach (1985)

Bluegill

96-h acute

LC50  11 ppbj

Gingerich and Rach (1985)

Snail (Physa pomilia)

24-h acute

LC50  6.35 ppm

Ling (2003)

a

Recorded after 15 days. Recorded after 7 days. c Recorded after 24 h. d Minimal lethal dose. e In hard water. f In soft water. g Determined at 20–23°C water temperature. h Maximum nonlethal concentration. i Mortality was 100% within 1 h; lower concentrations were not tested. j Determined at 13°C water temperature. b

Large doses of derris elicited convulsion in rabbits (Ambrose and Haag, 1936). When inhaled, derris was more toxic than pure rotenone, indicating high toxicity for other plant constituents (Ambrose and Haag, 1936). Derris and rotenone do not seem to be toxic to birds. For 12 different nestling birds, including chicken, pigeon, lark, sparrow, and pheasant, the oral LD50 values ranged from 0.1 to 0.3 g/kg (Cutkomp, 1943). The toxicity of rotenone used as a piscicide is influenced by environmental factors such as water turbidity, temperature, pH, and dissolved oxygen. It is most effective in acidic, clear waters that have little aquatic vegetation. Exposure time also affects toxicity. For example, the exposure time required for death at 50 ppb rotenone concentration was 3 h for yellow perch and bluegill but 11.25 h for the more resistant common carp (Rach and Gingerich, 1986). According to Willis and Ling (2000), exposing mosquitofish to 56 and 158 g/l rotenone, the toxic symptoms (color changes, surface air-gasping) occurred for 50% of the fish after 158 and 26 min, respectively, while 90% of mosquitofish were affected after 246 and 36 min, respectively.

In this study, black mudfish were approximately twice as sensitive as mosquitofish to rotenone. Upon transfer to fresh water, essentially all intoxicated individuals recovered. To kill fish in lakes and reservoirs, rotenone is applied at low rates; thus, consumption of poisoned fish by either wildlife or humans is not dangerous. Irritation  Derris powder was shown to be a mild local dermal irritant. However, it produced intense, though reversible irritation to the rabbit eye as well as severe pulmonary irritation to animals when inhaled (Ambrose and Haag, 1936; Haag, 1931). Rotenone itself is less irritating. In humans, it produces a sensation of numbness when applied over the mucous membrane of the mouth (Haag, 1931). Poisoning symptoms  Depending on the route of administration, the poisoning symptoms appear between 2 and 20 min and include initial increased respiratory and cardiac rates, incoordination, clonic and tonic spasms, and muscular depression. Temperature changes are not characteristic.

Chapter | 3  Pest Control Agents from Natural Products

Later, respiration slows down until it ceases but the heart continues to beat (Haag, 1931; Santi and Tóth, 1965). Death generally occurs in a few minutes after intravenous administration and within 2 days after intraperitoneal, subcutaneous and oral treatments. Continuous application of sublethal doses can also lead to mortality after several days of treatment. Typical symptoms of rotenone poisoning in the common carp were observed by Fajt and Grizzle (1998). Within minutes, carp increased respiration rates and activity, such as rising to the surface and gulping air, and displayed other behavioral changes. Blood oxygen concentration, first only in arteries and after a short lag time also in veins, increased up to 10-fold, whereas CO2 concentration was unchanged. Plasma pH decreased significantly, indicating a switch from aerobic to anaerobic metabolism. Death ensued after 35–40 min at 0.1 ppm rotenone concentration. Chronic toxicity  In a 37-day toxicity study with rats, 5 or 10 mg/kg daily oral doses of rotenone decreased food intake and reduced the weight and the survival rate of the animals as compared to the control animals (Lightbody and Mathews, 1936). In a 70-week study with rats, an emulsifiable concentration, containing 2.5% rotenone and 2.5% sulfoxide synergist, at 100 ppm in the drinking water showed no ill effect (Brooks and Price, 1961). In a 6-week trial, rabbits receiving food with 150 mg/kg rotenone showed no poisoning symptoms (Haag, 1931). For rotenone at 7.5, 37.5, or 75 ppm concentration in food fed to rats and at 0.4, 2, or 10 mg/kg given orally to beagles for 26 weeks, lower food intake and lower body weight gains were noted at the highest doses only (California EPA, 1997). In a rodent model for parkinsonism, daily intranasal application of 2.5 mg/kg rotenone for 30 days did not produce any noticeable motor alterations or damage to the nigrostriatal system (Rojo et al., 2007). Pathology  Dogs receiving rotenone at a chronic daily dietary dose of 10 mg/kg for a month displayed toxemia chiefly manifest in hepatic fatty metamorphosis and onethird of the bulk of liver occupied by fat (Haag, 1931). (For additional neuropathological symptoms, see above.) Effect on reproduction  Pregnant guinea pigs fed 150 ppm dietary rotenone had litters that were either born dead or died within 5 days after birth (Haag, 1931). The 75-ppm dietary concentration was tolerated better, but the weight gain of the surviving young was slower than for the control animals. Khera et al. (1982) found that administration of daily oral doses of 2.5 mg/kg cubé extract containing 87% rotenone to female rats on days 6–15 of gestation was without effect, but daily 5 mg/kg doses resulted in increased frequency of skeletal aberrations of the fetuses such as extra rib and delayed ossification of sternebrum.

141

The 10 mg/kg dose reduced maternal body weight, 60% of the dams were killed, and it also produced a high incidence of resorption in the surviving dams. A series of reproductive studies with 95–98% pure rotenone were summarized (California EPA, 1997). In a chronic feeding study with male and female rats, the effects of dietary rotenone on generations F0–F2, including breeding, gestation, lactation, and weaning, were observed. At 75 ppm, the highest concentration tested, decreased mean live litter sizes in the F0 and F1 generations were seen. At 37.5 and 75 ppm, decreased mean birth weights and decreased body weight gain of the F1 and F2 pups were observed. In a related 3-month one-generation study with Syrian golden hamsters, dietary rotenone at 1000 ppm concentration resulted in smaller testicles and infertility; at 500 ppm and above, poor litter survival, maternal death, and cannibalism were observed. In a teratology study with mice, involving rotenone at an oral dose of 24 mg/kg/day, given on days 6–17 of gestation, no maternal toxicity was observed, but decreased live litter size and increased fetal resorption were noted. Rotenone and other Complex I inhibitors inhibited ovulation in female rats (Koshida et al., 1987) as well as motility of rat and human sperms in vitro (Hong et al., 1983; Koppers et al., 2008). Treatment  Extending the observations of Ernster et al. (1963) (see also Wijburg et al., 1990) that inhibition by rotenone of rat liver mitochondrial respiration in vitro could be overcome by the addition of catalytic menadione (vitamin K3), Santi and Tóth (1965) demonstrated that an intravenous dose of 2.5 mg/kg menadione could rescue a rabbit from a 0.4-mg/kg lethal intravenous dose of rotenone (see also Tóth et al., 1966). (e)  Human Poisoning Incidents Acute rotenone poisoning is rare. De Wilde et al. (1986) described a fatal case of a 3.5-year-old girl who accidentally ingested approximately 0.6 g rotenone in about 10 ml of mixed ethereal oil insecticide formulation. The estimated oral dose was 40 mg/kg. Half an hour after ingestion, the girl vomited and felt drowsy. The initial symptoms rapidly developed to slow and irregular respiration, coma, and apnea. In spite of artificial respiration and gastric lavage, the girl died of respiratory arrest 8.5 h after ingestion. Significant postmortem findings included anoxic hemorrhages in the lung, heart, and thymus; anoxic damage to the cerebrum; bloody stomach content; and renal tubular necrosis, although the latter was suggested to be caused by the ethereal oils in the formulation. Analysis for rotenone in various tissues showed 1260 ppm in the stomach content, and 2–4 ppm in the blood, liver, and kidney but none in the brain, muscle, and thymus. It is noteworthy that the label on the insecticide, recommended for plants and external use on animals, stated “Natural Product – Non Toxic.”

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142

Rotenone was once also a common means of suicide in some parts of New Guinea. Victims taking small dosages usually recovered after gastric lavage and stimulants. For fatal cases, vomiting before death was usual and autopsy revealed acute congestive heart failure (Holland, 1938). Symptoms of a recent fatality following deliberate ingestion of approximately 200 ml of a commercial liquid derris formulation containing 0.8% rotenone have been recently described (Wood et al., 2005). The estimated amount ingested was 1.6 g rotenone, corresponding to a dose of 25 mg/kg. The typical poisoning symptoms were: initial vomiting followed by coma; blood pressure of 93/52 mmHg, heart rate 87 beats/min, liver dysfunction, and metabolic acidosis. Despite intubation and meticulous supporting care (norepinephrine infusion followed by dobutamine; empirical use of N-acetylcysteine, oral multivitamins, oral zinc sulfate, and intravenous iron as antioxidant therapy), the victim suffered an asystolic cardiac arrest and died 48 h after admission. Postmortem showed multiorgan failure, with pulmonary edema and congestion of the heart, spleen, and kidneys. The liver was icteric with centrilobular necrosis and general disintegration.

3.2.1.5  Sabadilla Alkaloids (a)  Introduction Liliaceous plants are known both for their poisonous properties and for their medicinal value. The crushed seeds of Schoenocaulon officinale (earlier Veratrum sabadilla), indigenous to Central America and the northern region of South America (Zomlefer et al., 2006), are known as sabadilla powder and were used by the American Indians in pre-Columbian times as an insecticide (e.g. louse powder, vermifuge as “pulvis capucinorum” in Europe) (Crosby, 1971). The seeds contain 2–4% of a biologically active alkaloid mixture, called veratrine. Veratrine was first isolated by Pelletier and Caventou in 1819, and since then it has been the subject of extensive chemical and pharmacological studies. Vigorous alkaline hydrolysis of veratrine affords the steroid alkamine cevine, the structure of which was elucidated in 1954 (Barton et al., 1954). Veratrine is an approximately 2:1 mixture of cevadine and veratridine (Figure 3.4), which are the respective angeloyl and veratroyl esters of the essentially inactive veracevine (i.e. the 3-epimer of cevine), obtainable from veratrine by mild hydrolysis. Minute amounts of additional alkaloids are also present in the seed extract. Sabadilla alkaloids belong to the ceveratrum group of Veratrum alkaloids characterized by a C-nor-D-homosteroid skeleton containing 6–8 hydroxyl groups of which at least one is esterified (Kupchan and Flacke, 1967). Alkaloids of the jerveratrum group contain only 1–3 oxygens and occur as free alkamines or as glycosides in plants. (b)  Identity and Physicochemical Properties Cevadine  IUPAC name: 4,12,14,16,17,20-hexahydroxy-4,9-epoxycevan-3-yl [(Z)-2-methylbut-2-enoate].

N

H

OH H

H

R

O

3β O OH

OH

OH

OH

OH

H

veracevine

R=H

cevadine

R = (z)-CH3CH=C(CH3) CO

veratridine

R = 3,4-(CH3O)2PhCO

cevacine

R = CH3CO

3-O-vanilloylveracevine

R = 3-CH3O-4-OH-PhCO

Figure 3.4  Structures of sabadilla alkaloids.

Chemical Abstract name: (3,4,16)-4,9-epoxycevane3,4,12,14,16,17,20-heptol 3-[(2Z)-methyl-2-butenoate]. CAS Registry Number: [62-59-9]. Empirical formula: C32H49NO9; molecular weight: 591.7. The melting point of cevadine is 208–210°C (decomposition); the optical rotation is [ ]20 D  10.7 (c  6.0 in ethanol). Veratridine  IUPAC name: 4,12,14,16,17,20-hexahydroxy-4,9-epoxycevan-3-yl 3,4-dimethoxybenzoate. Chemical Abstract name: [3,4,16]-4,9-epoxycevane3,4,12,14,16,17,20-heptol 3-(3,4-dimethoxybenzoate). CAS Registry Number: [71-62-5]. Empirical formula: C36H51NO11; molecular weight: 673.8. The melting point of veratridine is 167–184°C (decomposition); the optical rotation is [ ]20 D  7.2 (c  3.9 in ethanol). Veratridine is a weak base with a pKa of 9.54. The solubility of veratridine in a 150-mM aqueous NaCl solution is 12.5 g/l at pH  8.07. Sabadilla alkaloids are freely soluble in dilute acids but decompose in solutions with a pH greater than 10. They are readily soluble in alcohols, ether, and chloroform but not in hexane. The CAS Registry Number of sabadilla or the veratrine mixture is [8051-02-3]; CAS Registry Numbers of minor sabadilla alkaloids: veracevine: [587623-3]; cevacine: [28111-33-3]; sabadine: [124-80-1]; 3-Ovanilloylveracevine: [187237-90-7]. (c)  History, Formulations, and Uses Powdered rhizomes of the related white hellebore (Veratrum album), growing in Europe and Asia, and the “green (or false) hellebore” or Indian poke (V. viride), indigenous to the eastern part of North America, were used

Chapter | 3  Pest Control Agents from Natural Products

to cure herpes, toothache, rheumatism, and catarrh, and drugs from these plants were also important hypotensive agents although of low therapeutic index (Kupchan and Flacke, 1967). Preparations from hellebore roots were once commercial insecticides against hemipteran and homopteran pests of fruits and vegetables but are not used any longer (Shepard, 1951). Due to its low persistence and compatibility with beneficial insects (Bellows and Morse, 1993), sabadilla reappeared in the late 1970s. It is formulated as dust, wettable powder, or water-soluble concentrate, which might contain sugar as an insect feeding stimulant, with 0.2–25% alkaloid content. Sabadilla is now used against thrips in citrus and avocado (Hare and Morse, 1997; Humeres and Morse, 2006) and in organic farming (Zehnder et al., 2007). Typical use rates are 20–100 g total alkaloid/ha. Dilute sabadilla preparations are also available in homeopathic medicines. Stability  When exposed to air and sunlight, sabadilla formulations rapidly lose activity requiring proper timing of application or frequent treatments. Field trials in a citrus plantation showed that the alkaloid level on leaves declined to 60% of the initial deposit within 20 h after spraying. Veratridine persisted somewhat longer than cevadine; nevertheless, degradation and rainfall resulted in no residual toxicity to citrus thrips 7 days after treatment (Hare and Morse, 1997; see also Yee et al., 2001). (d)  Biological Properties Mode of action  Cevadine, veratridine, and related lipophilic ceveratrum alkaloids cause activation of the voltage-sensitive Na channels of nerve, heart, and skeletal muscle cell membranes similar to pyrethrins (see above). Both veratridine and cevadine alter the ion selectivity of Na channels and cause persistent activation. The receptor for these alkaloids has not been isolated, but experiments indicate it is distinct from that of pyrethrin. Sabadilla alkaloids share a binding site with the botanical steroid alkaloid aconitine, the frog steroid alkaloid batrachotoxin, and the botanical diterpenoid grayanotoxins (reviewed by Bloomquist, 1996; Catterall et al., 2007; Honerjäger, 1982; Wang and Wang, 2003). Structurally, veratridine and cevadine differ only in their acyl group (R in Figure 3.4) and this difference is enough to cause quantitative and qualitative variations in insecticidal activity (Allen et al., 1945; Ikawa et al., 1945; Ujváry et al., 1991), in mammalian toxicity (Swiss and Bauer, 1951; Ujváry et al., 1991) as well as in their pharmacological (Honerjäger et al., 1992; Mendez and Montes, 1943) and electrophysiological (Leibowitz et al., 1987; Nánási et al., 1990; Ohta et al., 1973; Shanes, 1952) properties. Physiological and pharmacological activities  Concoctions from Schoenocaulon, Veratrum, and Zigadenus genera

143

have been used for centuries in the treatment of fever and circulatory disorders, and their pharmacology has been thoroughly investigated (Honerjäger, 1982; Krayer and Meilman, 1977; Kupchan and Flacke, 1967). Purified veratridine, which has replaced veratrine mixture, is a commonly used and well characterized pharmacological tool in various (electro)physiological studies. The ceveratrum alkaloids have a characteristic hypotensive effect not directly involving the CNS. They slow the heart and lower arterial blood pressure by reflexly stimulating medullary vasomotor centers without decreasing cardiac output (Bezold–Jarisch effect). The electrophysiological aspects of the cardiac activity of veratridine and related steroid alkaloids were discussed in detail (Honerjäger, 1982; Honerjäger et al., 1982). The alkaloids also affect respiration: low doses induce tachypnea and higher doses cause respiratory depression and apnea. Ceveratrum alkaloids increase salivation, strongly irritate mucous membranes and induce sneezing, and have a reflex emetic effect in mammals able to vomit (see, for example, Andrews et al., 2000; Swiss, 1952). The low therapeutic index of the ceveratrum alkaloids limits their medical use. The acute hypotensive action of Veratrum alkaloids can be blocked by barbiturates and the reflex action is counteracted by atropine (Frey and Weigmann, 1943). Local anesthetics, including cocaine (Matthews and Collins, 1983; Zimányi et al., 1989; see also Ragsdale et al., 1994) and procaine (Biró and Gábor, 1969; Nishizawa et al., 1988; Ohta et al., 1973) antagonize veratrine action in vitro. Nefopam, an analgesic with skeletal muscle relaxant activity, provided neuroprotection from veratridine-induced excitotoxicity in cultured neurons in vitro; when administered intraperitoneally to mice at a 25-mg/kg dose, it also prevented convulsions induced either by electroshock or isoniazid (Novelli et al., 2007). Because ischemic conditions produce simultaneous Na and Ca2 influx through cardiac Na channels, veratridine-induced intoxication has been suggested as an experimental model of ischemia in animals (see Wermelskirchen et al., 1991). Lakics et al. (1995) showed that cell death evoked by veratridine in rat cerebrocortical cell cultures could be inhibited by submicromolar concentrations of the cerebroprotective agent vinpocetine more effectively than by the prototype Na channel blocker anticonvulsant phenytoin. Moreover, larger doses of vinpocetine as well as tetrodotoxin inhibited veratridine-induced increase in Ca2 levels in rat hippocampal pyramidal cells (Zelles et al., 2001). The Ca2 entry blocker flunarizine also provided protection against neuronal damage induced by veratridine (Pauwels et al., 1989). Acute administration of n-3 polyunsaturated fatty acids was recently shown to reduce both peak and late Na currents induced by veratridine in cloned human cardiac Na channels presumably through stabilizing interactions with the inactivated state of the channel (Pignier et al., 2007).

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Silva Freitas et al. (2001) examined the effect of intramuscular injection of veratrine on muscle morphology and energy metabolism of Nile tilapia, a freshwater fish. The alkaloid mixture caused transitory focal vacuolation in muscle fibers and prominent interstitial edema in affected muscle tissues. Veratrine decreased the aerobic metabolism in oxidative (red) fibers, while in oxidative-glycolytic (white) fibers it had an opposite effect. Degeneration of mitochondria was also evident. Prior treatment with the Na channel blocker tetrodotoxin prevented the ultrastructural changes caused by veratrine indicating the involvement of ion channel activation and/or delayed inactivation. In a related series of experiments with rodents, veratrine caused short-lived degeneration of muscle mitochondria (diameter alterations, cristae disorganization, and rupture) both in vitro and in vivo (Silva Freitas et al., 2006). In studies in vitro, decreased oxygen consumption, inhibition of NADH, succinic dehydrogenase and cyclooxygenase activities were also found that suggested nonspecific membrane disruption of muscle mitochondria unrelated to Na channel activation. Such disturbances, however, were not seen for liver mitochondria. In addition to the studies mentioned above, a broad range of pharmacological and biochemical effects, including the release of various neurotransmitters such as ACh, norepinephrine, dopamine, and GABA, were shown to be triggered by the veratridine-elicited increase in Na permeability of neural cells in vitro (see Cunningham and Neal, 1981). Metabolism and excretion  Veratrum alkaloids are easily absorbed through the skin and mucous membranes and upon ingestion. The metabolism and excretion of veratridine and cevadine have not been studied. On oral administration to humans of a structurally related steroid alkaloid (3-Oacetylgermine), slow urinary excretions of both the parent ester and its hydrolytic product, germine, were observed (Büch, 1971). In insects, the toxicity of sabadilla alkaloids can be synergized by PB, indicating oxidative detoxification (Blum and Kearns, 1956; Ujváry et al., 1991), although contribution of esterase inhibition by PB (see Khot et al., 2008), blocking the hydrolysis of veratridine and cevadine into the inactive veracevine, cannot be excluded. Recent pharmacokinetic studies by Li et al. (2009) with protoveratrine A administered to rats either intravenously (12.5 g/kg) or intragastrically (400 g/kg) showed essentially identical plasma half-lives (about 22 h) of the alkaloid. (e)  Toxicity to Laboratory Animals Acute toxicity  Acute toxicity data for the natural sabadilla alkaloids are listed in Table 3.4. Intravenously administered veratridine kills mice usually within 4 min; on intraperitoneal injection, the alkaloid causes death within 1.5 h after injection due to respiratory failure. Salivation and cyanosis are also notable and a short period of convulsion can

precede death. The poisoning symptoms of other Veratrum alkaloids also unfold almost immediately, but paralysis followed by severe convulsions could occur (Krayer et al., 1944; Mendez and Montes, 1943). The intraperitoneal LD50 value for protoveratrine was 0.44 mg/kg to the mouse (Swiss and Bauer, 1951). Teratogenicity  The ceveratrum-type sabadilla alkaloids do not appear to be teratogenic (California EPA, 2001). However, some members of the structurally distinct jerveratrum alkaloids, for example cyclopamine and jervine, are teratogenic. The bizarre birth defects (holoprosencephaly, including cyclopia) in livestock grazing plants such as V. californicum (Keeler, 1986, 1988; see also Lee et al., 2003b) are due to the inhibition of cholesterol biosynthesis and transport in embryos (Cooper et al., 1998). Table 3.4  Acute Toxicity of Sabadilla Alkaloids Animal

Assay

LD50 (mg/kg)

References

Mouse

ip

7.5a

Swiss and Bauer (1951)

Mouse

ip

8.5b

Swiss and Bauer (1951)

Rat

ip

4.8

Mendez and Montes (1943)

ip

100

Ujváry et al. (1991)

Mouse

ip

3.5

Swiss and Bauer (1951)

Mouse

ip

5.8

Ujváry et al. (1991)

Rabbit

sc

0.5–1.3

Krayer et al. (1944)

Frog

sc

1.5–30

Krayer et al. (1944)

Rat

ip

3.5

Mendez and Montes (1943)

Mouse

ip

0.42

Krayer et al. (1944)

Mouse

ip

1.35

Swiss and Bauer (1951)

Mouse

ip

9.0

Ujváry et al. (1991)

ip

5c

Veratrine

Veracevine Mouse Cevadine

Veratridine

3-OVanilloylveracevine Mouse a

From Merck & Co. From S. B. Penick & Co. c No mortality at this exploratory dose (Ujváry and Casida, unpublished observations). b

Chapter | 3  Pest Control Agents from Natural Products

(f)  Toxicity to Humans Irritation  Sabadilla and related Veratrum alkaloids have a sternutatory action, a property that has caused several bizarre intoxications among children. Eye contact could result in severe irritation, lacrimation, and inflammation of the conjunctiva. A recent report summarized seven poisoning cases and warned of the potential danger of V. album-based sneezing powders marketed in some European countries in the 1980s (Carlier et al., 1983; Fogh et al., 1983). Poisoning incidents  Despite the long and widespread use of ceveratrum alkaloids for the treatment of tachycardia and various circulatory disorders, especially hypertension, no fatal poisoning has been documented (Krayer and Meilman, 1977; Rokin and Kustovskii, 1997). Documented sabadilla poisoning cases are also rare (see Ray, 1991). Because sabadilla alkaloids from Sch. officinale and the structurally somewhat different ceveratrum alkaloids of other Veratrum plants have a similar mode of action, poisoning cases reported for the latter could be instructive. Veratrum plants are sometimes mistaken for other edible or medicinal plants, often Gentiana lutea, causing intoxication (reviewed by Schep et al., 2006). Jaffe et al. (1990) described six poisoning cases due to accidental ingestion of V. viride. The typical signs of intoxication were abdominal pain, nausea, vomiting, diaphoresis, bradycardia, and hypotension developing from 0.5 to 4.5 h after ingestion. Similar V. album poisoning cases have been described (Festa et al., 1996; Zagler et al., 2005). Grobosch et al. (2008) identified protoveratrines but not jervine, veratridine or cevadine in the beverage and in the serum of a poisoned person who had ingested two glasses of self-made alcoholic beverage made from roots of V. album thought to be G. lutea. Gaillard and Pepin (2001) have described two European fatal poisoning cases believed to be caused by V. album ingestion. Postmortem inspection of the stomach content revealed seeds, which were curiously shown to contain cevadine and veratridine, previously reported solely from Sch. officinale, but no other alkaloids. The two sabadilla alkaloids were also present in the heart blood and gastric content of the deceased. V. album, however, is known to contain the toxic angeloyl and veratroyl esters of zygadenine, a structural isomer of veracevine differing from it by the relative positions of the hydroxyl groups on the epoxycevane skeleton (see Gfeller et al., 1995). Zygadenine derivatives are found in Zigadenus (death camas) and in some Veratrum species. Quatrehomme et al. (1993) surveyed 32 Veratrum poisoning cases, one of which was an accidental ingestion of veratrine antilouse preparation.

145

and vomiting have subsided. Subsequently, antiemetics can be used. Seizures, if they occur at all, may be treated with standard anticonvulsants. Bradycardia is responsive to intravenous atropine, and hypotension can be treated with dopamine or metaraminol. The patients usually recover within 48 h.

3.2.1.6  Ryania (a)  Introduction The botanical insecticide ryania is the ground stemwood of Ryania speciosa (Flacourtiaceae), a tropical tree growing in Central America and the Amazon Basin (Crosby, 1971). Related Ryania species in that region were used as the source of arrow poisons. Ryanodine was isolated from the roots and the stemwood of R. speciosa by Rogers et al. (1948), and the structure of the alkaloid was established by Wiesner (1972) as the pyrrolecarboxylate ester of the diterpene ryanodol (Figure 3.5). Another major insecticidal component of the wood is 9,21-didehydroryanodine (Waterhouse et al., 1984; see also Cabras et al., 2001). Thirty additional minor ryanoids have been identified from the plant (see Ruest et al., 2002). The total active alkaloid content of ryania insecticide is about 0.22%. Nonester ryanoids with insecticidal and/or antifeedant activity have also been isolated from Persea indica (González-Coloma et al., 1999). (b)  Identity, Physicochemical Properties, and Uses Chemical Abstract name of ryanodine: ryanodol 3-(1H-pyrrole-2-carboxylate). Chemical Abstract name of ryanodol: (3S,4R,4aR,6S,6aS, 7S,8R,8aS,8bR,9S,9aS)-hexahydro-3,6a,9-trimethyl-7(1-methylethyl)-6,9-methanobenzo[1,2]-pentaleno[1,6bc]furan-4,6,7,8,8a,8b,9a,(6aH,9H)-heptol. CAS Registry Numbers: ryania: [8047-13-0]; ryanodine: [15662-33-6]; 9,21-didehydroryanodine: [94513-550]; ryanodol [6688-49-9]; 9,21-didehydroryanodol: [106821-54-9]. Empirical formula of ryanodine: C25H35NO9; molecular weight: 493.6. HO

HO

H

OH HO

HO 3

21 9

O

RO OH ryanodol R = H ryanodine R =

Treatment  The treatments employed in Veratrum alkaloid poisoning cases are supportive (Jaffe et al., 1990). Charcoal and a cathartic can be administered after nausea

N H

O

Figure 3.5  Structures of ryanodine and ryanodol.

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Ryanodine melts with decomposition and the reported melting points are 219–220°C (Rogers et al., 1948) and 235–237°C (Waterhouse et al., 1984). Ryanodine is soluble in water, ethanol, acetone, ethyl ether and chloroform, but practically insoluble in hexane. Ryanodine is dextrorotatory: [ ]20 D  26 (c  1.02 in methanol).

Although the importance of ryania and ryanodine, as botanical insecticides, has diminished during the past decades, there is a renewed interest in RyRs due to the recent discovery of new synthetic compounds, exemplified by flubendiamide and rynaxypyr, targeting this site (Nauen, 2006; see also elsewhere in this book).

Stability  Ryania is relatively stable on exposure to light and air and has a longer residual activity in the field than rotenone and the pyrethrins but no adequate evaluation of the stability and environmental fate of ryania or its constituents has been carried out.

Biochemical effects and pharmacology  In the cat, intraarterial injection of 50 g of the alkaloid induced tetanic contractures in skeletal (tibialis anticus) muscle, which could be abolished by intraarterial injection of 1 mg atropine (Procita, 1956). At nanomolar concentrations, ryanodine caused an irreversible loss of contractile tension in cat cardiac muscle that could be temporarily reversed by sympathomimetics or excess calcium (Hillyard and Procita, 1959). In a related study with dogs, Procita (1958) found that injection of lethal doses (50–300 g/kg) of ryanodine produced enophthalmos, followed by general spastic muscle rigidity, salivation, vomiting, and defecation. Spastic rigidity then progressed and death was due to asphyxiation, although in some cases cardiac arrest could also be observed. A somewhat different picture emerged when animals were anesthetized with pentobarbital: A single lethal dose of ryanodine, without affecting other muscles, produced circulatory failure before respiratory difficulties arose and artificial respiration was of no value. The hypotension produced by sublethal doses of ryanodine was unaffected by atropine and cardiac glycosides but epinephrine reversed normal circulation. The potency of ryanodine (IC50  10 nM) and analogs in blocking the rabbit skeletal muscle RyR proved to be a good indicator of their toxicity to mice, establishing the toxicological relevance of these assays (Jefferies et al., 1992; Pessah et al., 1985; Waterhouse et al., 1987; see also Sutko et al., 1997). In nonmammalian vertebrates, in particular the fish, the influx of extracellular Ca2 through voltage-gated calcium channels also contributes to the regulation of cytosolic Ca2 content. Temperature changes (see, for example, Rocha et al., 2007; Vornanen, 2006) as well as training (see, for example, Anttila et al., 2008) can shift muscle contraction control from a mainly ion channel-dependent one to a mainly RyR-dependent one (or vica versa) within an individual. Thus, isolated heart muscle preparations of the Pacific mackerel were more sensitive to ryanodine at 20°C than at 15°C, indicating that at lower temperatures calcium ion channels play a predominant role in atrial muscle contractility in this fish (Shiels and Farrell, 2000). Moreover, the effects of ryanodine could be ameliated by epinephrine in this study. Other fish, particularly species with active lifestyles in the cold, however, are generally sensitive to ryanodine at low temperatures (see Tiitu and Vornanen, 2003 and references therein). The binding of [3H]ryanodine to SR RyR-A preparations from blue marlin and tuna fast- and slow-muscles displays classical

Formulations and uses  Commercial ryania is formulated as a dust or water-dispersible powder. Ryania is a slow-acting contact and stomach insecticide and is often mixed with other botanical preparations, but its use is now restricted. The usual application rates of ryania are 10–72 kg/ha (20–145 g alkaloid equivalent/ha) against lepidopteran pests (Crosby, 1971). (c)  Biological Properties Mode of action  Reviews on ryanodine pharmacology in mammals (Sutko et al., 1997) and in insects (Jefferies and Casida, 1994) as well as on receptor structure and function (Coronado et al., 1994; Sattelle et al., 2008) are available. Ryanodine acts by binding to a family of intracellular Ca2 release channel proteins, the so-called ryanodine receptors (RyRs), associated mainly with the sarcoplasmic reticulum (SR) of skeletal and cardiac muscles but also detected in the brain and liver. The alkaloid at submicromolar concentrations locks the Ca2 channels in a sustained, though fractional conductance state, whereas at higher concentrations the channel is transformed to a nonconducting, closed state. The outflow of Ca2 from SR through the open channel into the cytosol of muscle cells results in muscle contraction. The RyRs are composed of large polypeptides with a molecular weight of 550–565 kDa forming a mushroomshaped ion channel in a homotetrameric arrangement (Takeshima et al., 1989). Different RyR isoforms were isolated from various organisms and, in some cases, also from different cell types of the same organism. In mammals, there are three ryanodine receptor types: RyR1 (mainly in skeletal muscle), RyR2 (mainly in cardiac muscle), and RyR3 (at low levels in various muscle and other tissues, including brain). Avian, amphibian and fish skeletal muscle, however, possess only two isoforms, RyR-A and RyR-B resembling mammalian RyR1 and RyR2, respectively, and these isoforms could be detected together in the same muscle fiber. Insect RyRs are less characterized. The amino acid sequence identity of known insect RyRs is 77% but these receptors show only approximately 47% sequence similarity with the known mammalian RyRs (see Sattelle et al., 2008).

Chapter | 3  Pest Control Agents from Natural Products

bell-shaped Ca2 dependency similar to mammalian skeletal muscle RyR1 (activation at micromolar Ca2concentration and inactivation at millimolar Ca2-concentration), and the slow-twitch muscle RyR-A has a greater sensitivity for Ca2; receptor protein analysis also indicates the existence of fiber type-specific receptor isoforms (Franck et al., 1998). The complexity of the pharmacological effects of ryanodine is further illustrated by its distinct paralyzing action in different striated muscles. The alkaloid induces flaccid paralysis in cardiac muscles but rigid paralysis in skeletal muscle of vertebrates. In cardiac muscle, ryanodine elicits excessive release of Ca2 from intracellular stores by locking the RyR2 channels in the long-lived subconducting state. This “leaked” Ca2 is then rapidly removed from the cell by an effective surface membrane Ca2 extrusion mechanism present in cardiac muscle. The flaccid paralysis is eventually due to the depletion of intracellular Ca2 stores. By contrast, the SR of skeletal muscle lacks this Ca2 extrusion mechanism so ryanodine-induced Ca2 leak from intracellular stores leads to Ca2 accumulation in the cytoplasm. The resulting abnormally high cytoplasmic Ca2 level induces sustained skeletal muscle contraction, that is rigid paralysis (for a detailed discussion, see Fill and Copello, 2002). In addition to ryania alkaloids, the RyRs are modulated by other agents, including Ca2, nitric oxide, adenine nucleotides, caffeine, calmodulin, digoxin, general (volatile) and local anesthetics, the muscle relaxant dantrolene, verapamil, and some immunosuppressant macrolides (Zucchi and Ronca-Testoni, 1997), as well as by polyhalogenated hydrocarbons (see, for example, Pessah et al., 2009). For example, millimolar concentrations of Mg2 inhibited, whereas caffeine stimulated [3H]ryanodine binding in vitro to skeletal receptors while cardiac receptors were only slightly affected by these agents. The binding of [3H]ryanodine to skeletal and cardiac RyR was stimulated by increasing the pH, the temperature, or the NaCl or KCl concentrations of the preparation (see Coronado et al., 1994). Abnormal functioning of RyRs is implicated in various myocardial diseases, malignant hyperthermia and some skeletal muscle diseases, as well as in cancer. Metabolism and excretion  Except for one report showing no detectable insecticide in the urine of rats receiving 600 mg/kg ryania powder orally (Kuna and Heal, 1948), the fate of ryanodine and other ryania alkaloids in mammals is not known. (d)  Toxicity to Laboratory Animals Acute toxicity  The acute toxicity of ryania powder and pure ryanodine to various animals is shown in Table 3.5. Ryanodol, the hydrolysis product of ryanodine, has relatively low toxicity to mice (intraperitoneal LD50  20 mg/kg) and little activity at the mammalian SR RyR

147

(IC50  35 M), yet it is a potent insecticide (Waterhouse et al., 1987), demonstrating that selective toxicity could be achieved among ryania alkaloids. Whether the selectivity is due to species-dependent differences in distribution, detoxification, or properties of the Ca2 channel is not clear (see Jefferies et al., 1997 and references therein). It was also demonstrated that both ryanodine and the nonester 9,21didehydroryanodol, which is insecticidal but has low mammalian toxicity to mice (intraperitoneal LD50  20 mg/kg), affected the excitability and ion selectivity of K channels in insects but not in mammals (Usherwood and Vais, 1995; Vais et al., 1996). Chronic toxicity and pathology  Kuna and Heal (1948) studied the chronic toxic effects of the powdered stemwood of R. speciosa on various animal species. Rats, guinea pigs, and chickens remained symptomless for 5 months when fed a diet containing 1% ryania powder. Rats fed 5% ryania showed decreased weight gain, and some deaths occurred within 25 days after the start of the treatment. Rats exposed to ryania dust or a spray of 1% aqueous ryania suspension for 8 h daily during 22 days did not display any treatment-related toxic symptoms. Dogs, however, showed signs of toxicity (vomiting, irritated eyes and respiratory passages) when exposed to ryania powder for 2 h. Autopsy of rats receiving 2–5% ryania powder in the diet revealed hemorrhages in the pancreas and intestinal tract, pulmonary complications, and pleural exudation. Treatment  Based on the similarities between some of the effects of ryanodine and malignant hyperthermia, a life-threatening genetic disorder of skeletal muscle characterized by uncontrolled Ca2 release involving RyRs, ryanodine toxicity is considered to be a model of this rare disease. Accordingly, dantrolene, which is used for the clinical treatment of malignant hyperthermia, protects mice and rats from lethal intraperitoneal doses (135 g/kg) of ryanodine (Fairhurst et al., 1980; see also Fruen et al., 1997; Paul-Pletzer et al., 2001). There appears to be no described animal or human poisoning cases.

3.2.1.7  Azadirachtin (a) Introduction The neem tree, Azadirachta indica (Meliaceae), also known as “nim” or “margosa,” is indigenous to the arid parts of India and Burma (now Myanmar) and is now grown in Africa and other tropical and subtropical regions, including plantations in Thailand, southwest China, and Australia. Indian folklore and medicinal literature, including “Ayurveda,” consider it a miracle tree and the medicinal properties of neem are still of special interest (Brahmachari, 2004; Randhawa and Parmar, 1993; Singh and Singh, 2002). Traditional neem preparations from all

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148

Table 3.5  Acute Toxicity of Ryania and Ryanodine Animal

Assay

LD50 (mg/kg)

Other data

References

Rat

oral

1200

Kuna and Heal (1948)

Mouse

oral

650

Kuna and Heal (1948)

Guinea pig

oral

2500

Kuna and Heal (1948)

Rabbit

oral

650

Kuna and Heal (1948)

Dog

oral

150

Kuna and Heal (1948)

Monkey

oral

400

Kuna and Heal (1948)

Chicken

oral

3000

Ryania

Bobwhite quail

dietary

Kuna and Heal (1948) a

US EPA (1999)

a

NOEL  3160 ppm

Mallard duck

dietary

NOEL  5620 ppm

US EPA (1999)

Daphnia magna

48-h acute

44 ppm

US EPA (1999)

Ryanodine Mouse

ip

0.10

Rat

ip

0.32

Mouse

ip

0.26

Guinea pig

ip

0.21

Waterhouse et al. (1987) b

Procita (1958) Procita (1958)

b

Procita (1958) b

Procita (1958) Procita (1958)

Rabbit

iv

0.026

Cat

ip

0.071b

Dog

oral

Dog

iv

0.075

LD  0.4 mg/kgc

Procita (1958)

LD  0.1 mg/kgc

Procita (1958)

a

No observable effect level. Calculated from mortality data. c Lethal dose. b

parts of the tree, including the bark, roots, flowers, and seeds, have been used for centuries for medical, agricultural (pest control agent and fertilizer), hygienic, and cosmetic purposes. Neem seeds can be pressed to give 20–40% neem oil, a dark-yellow liquid of disagreeable, garlic-like odor. The oil has a bitter taste due to the presence of some 2% limonoid constituents. Azadirachtin (sometimes called azadirachtin A) (Figure 3.6) was isolated as one of the major bioactive limonoids from the seeds of A. indica (Butterworth and Morgan, 1968). Dried seed kernels may contain up to 0.9% azadirachtin. The structure of the highly oxygenated tetranortriterpenoid was elucidated fully in 1987 by several research groups (reviewed by Morgan, 2009; Veitch et al., 2008). 3-Tigloylazadirachtol (with 1-OH unesterified) and salannin (Figure 3.6) are also abundant in neem seed extracts. These and related neem triterpenoids are biosynthetically derived from the steroidal tirucallol by successive rearrangement and oxidation reactions. The chemistry of

azadirachtin and related limonoids has been reviewed (Akhila and Rani, 1999; Ley et al., 1993). (b)  Identity, Physicochemical Properties, and Uses IUPAC name: dimethyl (2aR,3S,4S,4aR,5S,7aS,8S, 10R,10aS,10bR)-10-acetoxy-3,5-dihydroxy-4[(1aR,2S,3aS,6aS,7S,7aS)-6a-hydroxy-7a-methyl-3a,6a,7,7atetrahydro-2,7-methanofuro[2,3-b]oxireno[e]oxepin-1a (2H)-yl]-4-methyl-8-{[(2E)-2-methylbut-2enoyl]oxy}octahydro-1 H-naphtho[1,8 a-c:4,5bc]difuran-5,10a(8H)-dicarboxylate. Chemical Abstract name: (2aR,3S,4S,4aR,5S,7aS,8S,10R, 10aS,10bR)-10-(acetyloxy)octahydro-3,5-dihydroxy4-methyl-8-[[(2E)-2-methyl-1-oxo-2-butenyl]oxy]4-(1aR,2S,3aS,6aS,7S,7aS)-(3a,6a,7,7a-tetrahydro-6a-hydroxy-7a-methyl-2,7-methanofuro[2,3b]oxireno[e]oxepin-1a(2H)-yl)-1H,7H-naphtho[1,8bc:4,4a-c]difuran-5,10a(8H)-dicarboxylic acid dimethyl ester.

Chapter | 3  Pest Control Agents from Natural Products

O 3'

2'

O

O

O

O

O

OH O

1

O 3

H

O

O

Figure 3.6  Structures of azadirachtin and other limonoids from the neem tree.

O O

OH 22 O

O

149

23 H

O O

OH

H

O

O

O

O

O

O nimbin

azadirachtin O

O

O

O

O O

O

O O

O O

O H

O H

O

salannin

O

Note: In practice, and throughout this chapter, the limonoid skeleton is numbered according to the steroid numbering system to reflect its biogenesis. CAS Registry Number: [11141-17-6]. Empirical formula: C35H44O16; molecular weight: 720.7. CAS Registry Numbers of some related compounds: 22,23-dihydroazadirachtin: [108189-58-8]; 2,3,22,23tetrahydroazadirachtin: [108168-76-9]; nimbin: [594586-8]; nimbolide: [25990-37-8]; salannin: [992-20-1]; 3-tigloylazadirachtol (azadirachtin B) [95507-03-2]. Physicochemical properties  For pure azadirachtin, the melting point is 160°C; at ambient temperature, azadirachtin is slightly soluble in water (1–3 g/l), readily soluble polar organic solvents (ethanol, ethyl acetate, acetone and chloroform) but insoluble in hexane. Its log P value is 1.09. Azadirachtin is levorotatory: []D  53° (c  0.5 in chloroform). Stability  Azadirachtin is relatively stable in crystalline form if stored in the dark. Its laboratory half-life in mildly acidic solutions (pH 4–6) is 50–100 days at room temperature, but rapid decomposition (hydrolysis, isomerization, and/or rearrangement) occurs at higher temperatures, in alkaline and strongly acidic media, and especially in the light (Barrek et al., 2004; Ermel et al., 1987; Jarvis et al., 1998; Sundaram et al., 1995; Szeto and Wan, 1996). The individual limonoids in commercial formulations degrade differently in the field as demonstrated in a recent experiment where the half-lives of azadirachtin, azadirachtin B,

O nimbolide

and salannin on strawberries were approximately 17, 23, and 1 h, respectively (Caboni et al., 2006). Neem formulations typically retain over 59% of their azadirachtin content for about a year when stored at 10–15°C in the dark. Interestingly, formulations that had lost 95% of azadirachtin upon storage at 54°C for 14 days were as insecticidal as unheated ones indicating the presence of other, more stable natural or transformed (artifact) products (Kumar and Parmar, 2000; see also Barnby et al., 1989a). Studies on the environmental fate of azadirachtin in various formulations have been reviewed (Stark, 2004; Sundaram, 1996). In the field, the residual life of regular azadirachtin extracts is 8–10 days; commercial formulations, however, usually contain stabilizers that retard both hydrolysis and photodegradation. The half-life of azadirachtin on foliage could be as short as 17 h, whereas in the soil, due to the absence of light, the half-life could be as high as 25 days. Similarly, for an azadirachtin formulation applied in a forest lake environment, the average time to 50% dissipation was about 26 days (Thompson et al., 2002b). Two stable hydrogenated derivatives of azadirachtin, namely 22,23-dihydro- and 2,3,22,23-tetrahydroazadirachtin (Barnby et al., 1989a; Dhingra et al., 2008; Immaraju et al., 1994; U.S. EPA, 1998a), have also been used in insect control. Formulations and uses  The most important neem products traditionally used in agriculture are aqueous neem

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seed kernel and leaf extracts, alcoholic seed kernel and leaf extracts, enriched and formulated seed kernel extracts, neem seed oil, and neem seed cake that remains from the kernels after pressing the oil therefrom. Neem seeds from different geographic regions vary considerably in their composition and the method of extraction also affects the limonoid content and the bioefficacy of the products (Ermel et al., 1987; Isman et al., 1990; Sidhu et al., 2003; Stark and Walter, 1995). The commercial and partially standardized formulations are based on refined extracts of neem seed kernel extracts or neem oil and sold as powder or emulsifiable liquid concentrates with specified (typically 0.1–25%) azadirachtin content. It should be pointed out again that the biological activity, including toxicity to nontarget organisms, of a formulation varies according to its azadirachtin content, the nature and the relative amount of other neem constituents might be different even for batches from the same manufacturer (see, for example, Goktepe and Plhak, 2002; Kumar and Parmar, 2000). Although neem leaf extracts have been shown to inhibit aflatoxin production by Aspergillus fungi (see RazzaghiAbyaneh et al., 2007 and references therein), the seeds in humid areas can carry this mycotoxin-producing fungus (Jacobson, 1989; see also Hansen et al., 1994). Proper hygienic practices during harvesting, processing, handling (drying to moisture content below 14%), and storage minimize aflatoxin contamination. Azadirachtin-containing products are used against a broad range of insect pests in orchards, vegetables, mushrooms, herbs, tea, coffee, cotton, turf, and ornamentals as well as for disease vector control. Typical use rates are low (10–40 g/ha), but due to the instability of the active ingredient(s) frequent applications may be required (Immaraju, 1998). Due to its systemicity, azadirachtin can also be applied to the soil (see, for example, Thoeming et al., 2006). Neem extracts have other agriculturally important biological activities, including nematicidal (Alam, 1993) and antifungal (Coventry and Allan, 2001; Parveen and Alam, 1993) effects. Although neem oil and seed cakes have been used as fertilizers without noticeable problems, preparations rich in azadirachtin might disturb soil microflora if applied at high rates (Gopal et al., 2007). Recently, shampoo formulations of neem seed extracts have also been developed for veterinary and human use as antiparasitic agents (Abdel-Ghaffar et al., 2008; Heukelbach et al., 2006). An interesting property with toxicological relevance of neem leaves and stem bark powder is their capacity to adsorb heavy metals, for example Cd2 and Pb2 (Sharma and Bhattacharyya, 2005) as well as Zn2 (Arshad et al., 2008) from aqueous solution. (c)  Biological Properties Mode of insecticidal action  Neem extracts and azadirachtin are nonneurotoxic pest control agents exhibiting

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unparalleled insecticidal effects. Physiologically, they exhibit strong behavioral, growth regulatory and reproductive activities and the subject has been reviewed extensively (Ascher, 1993; Mordue, 2004; Mordue et al., 2005; Schmutterer, 1987, 1990). In spite of considerable research efforts, the mode of action of azadirachtin, which is the most significant neem component, has not been clarified at the cellular or biochemical level. Neem preparations and azadirachtin have deterrent or antifeedant activities against many insects, especially polyphagous species, and inhibit feeding at concentrations of 0.01–1 ppm. According to various electrophysiological studies, azadirachtin stimulates chemoreceptors on “deterrent” cells and blocks the firing of phagostimulatory (e.g. sucrose) receptors on feeding stimulatory cells (Koul, 2008). Azadirachtin markedly affects insect metamorphosis and reproduction, including fecundity, but these effects manifest slowly. Depending on the dose, azadirachtin causes growth inhibition, malformation, and mortality in insect larvae. It disturbs insect development apparently by interfering with the biosynthesis, release or action of ecdysteroids and/or other hormonal regulators of insect molt. For example, submillimolar concentrations of azadirachtin as well as salannin and nimbin inhibited ecdysone 20-monooxygenase, required for the biosynthesis of 20-hydroxyecdysone, the key insect molting hormone (Mitchell et al., 1997). Among a series of neem limonoids and synthetic analogs studied by Salehzadeh et al. (2002), azadirachtin, 22,23dihydroazadirachtin, 2,3,22,23-tetrahydroazadirachtin and azadirachtin B effectively inhibited replication of cultured S. frugiperda ovary cells at submicromolar concentrations. Nimbin, salannin and some minor neem terpenoids were less active. Furthermore, azadirachtin was up to a million times more toxic to insect cells than cells of murine fibroplast, human liver and breast tissue. Furthermore, azadirachtin irreversibly inhibited proliferation of various cultured insect cells, but not murine fibroplast cells, apparently by interfering with the formation and assembly of mitotic spindles in a manner similar to that of the antimitotic alkaloid colchicine; the limonoid also inhibited the polymerization of tubulin from pig brain but much less effectively than colchicine (Salehzadeh et al., 2003). Azadirachtin, at submicromolar concentrations, has been found to bind to a specific nuclear protein complex showing sequence similarity to heat shock (stress) proteins in Drosophila cell culture (Robertson et al., 2007). In addition, azadirachtin has been reported to bind to actin proteins from Drosophila and lepidopteran species in vitro and to lead to apoptosis in vivo (see Anuradha and Annadurai, 2008). Azadirachtin has also been shown to inhibit protein synthesis in various insect tissues including midgut and fat body cells and also to impair food protein digestion. In addition, in laboratory assays sublethal dietary pre-exposure

Chapter | 3  Pest Control Agents from Natural Products

to azadirachtin increased susceptibility of larvae of a lepidopteran species to carbamate, organophosphate and pyrethroid insecticides most likely by reducing enzymatic activity (detoxification) in azadirachtin-pretreated individuals (Lowery and Smirle, 2000). In summary, the various anti-insectan effects of azadirachtin are probably due to the disturbance of multiple cellular processes. Because commercial neem formulations contain other bioactive but less studied limonoids, the mode of toxic action of such preparations is obviously more complex than that observed for pure azadirachtin. Metabolism and excretion  Information on the fate of azadirachtin in animals and plants is scarce. The complex structure of azadirachtin and congeners prevented the chemical or biological characterization of the metabolites formed. Injection of a radiolabeled bioactive azadirachtin derivative, namely [22,23-3H2]dihydroazadirachtin into locusts indicated fast clearing from the blood. Ninety percent of the applied radioactivity was excreted with the feces during the first 7 h, whereas the rest of the material accumulated in the Malpighian tubules where it could be detected even 24 days after treatment. After the first 24 h, feces contained at least three polar, unidentified metabolites but no [22,23-3H2]dihydroazadirachtin (Rembold et al., 1984, 1988; see also Paranagama et al., 1993). However, slower excretion of the same compound was found upon oral application into larvae of Heliothis virescens (Barnby et al., 1989b) or Peridroma saucia (Koul et al., 1994). For both species, approximately 50% of the administered radioactivity was excreted by 72 h after treatment as a single polar dihydroazadirachtin metabolite and for the latter species the retained radioactivity was mostly found in the gut (24%) and the integument (12%). Effects of neem preparations on mammals  The pharmacological and toxicological properties of neem extracts, especially neem oil commercialized in India, have been studied extensively. The broad spectrum of activity of neem products includes anti-inflammatory, antipyretic, analgesic, cardiovascular, hypoglycemic, diuretic, immunomodulatory, dermatological, antimicrobial, antifungal, antimalarial, antiparasitic, and antifertility effects (reviewed by Dhawan and Patnaik, 1993; Kanungo, 1993; Subapriya and Nagini, 2005). Effects on reproduction  The effect of neem preparations on reproduction has received special interest (see Boeke et al., 2004b; Brahmachari, 2004). In general, among the possible adverse effects of neem used as a pesticide the most critical ones are reversible reproduction disturbances in both male and female mammals upon subacute and chronic exposure. However, one has to bear in mind that the exact chemical composition of the extracts used in the studies was unknown.

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Oral administration of neem seed extract to female rats from days 8 to 10 of pregnancy caused complete resorption of embryos by day 15 of pregnancy and the animals regained fertility in cycles subsequent to treatment (Mukherjee and Talwar, 1996). Single intrauterine administration of 100 l neem oil caused lasting infertility by apparent induction of leukocytic infiltration in the uterine epithelium during the preimplantation period (between days 3 and 5 postcoitum). Fertility was regained 5 months after treatment without apparent teratogenic effects (Upadhyay et al., 1990). Similarly, subchronic oral or dietary administration of a technical neem extract containing 12% azadirachtin to pregnant female rats did not cause significant embryotoxic or teratogenic effects and there were no treatment-related adverse effects on the reproductive performance of females of two generations under continuous treatment (Srivastava and Raizada, 2007). In one of the early studies (Sadre et al., 1984), daily oral doses of neem leaf extract to male mice and rats affected neither normal development nor spermatogenesis but caused reversible infertility, which was attributed to the decreased motility of spermatozoa. The extract, however, was toxic to guinea pigs and rabbits. By contrast, neem seed kernel extracts lacked activity in a similar test with rats (Krause and Adami, 1984). Daily intramuscular injection of 250 and 500 mg/kg doses of neem oil for 8 days to male rats caused significant decreases in sperm counts, epididymal weight, and glycogen levels; reduced acid phosphatase and influenced lactate dehydrogenase activities; and increased alkaline phosphatase activity. Marked structural changes in the testes and impaired spermatogenesis were also observed. It was suggested that neem oil impaired the androgen supply to the testicular and epididymal tissues (Manoranjitham et al., 1993; Sampathraj et al., 1993; see also Aladakatti and Ahamed, 2005; Ghodesawar et al., 2004). Investigations of the antifertility (spermicidal) property of neem oil in rodents and humans culminated in the commercialization of human contraceptive formulations in India (see Riar and Alam, 1993; Singh and Singh, 2002; Subapriya and Nagini, 2005; Talwar et al., 1997). (d)  Toxicity in Animals Acute and chronic toxicity  As with the pharmacological studies, toxicological evaluations rarely used pure ingredients but tested neem oils, extracts from various parts of the tree, or formulated insecticide products instead (Jacobson, 1989; Kanungo, 1993). Consequently, results from studies using different extracts containing several active ingredients in varying ratio could differ. The acute toxicity values of several neem preparations and pure azadirachtin for laboratory animals and some nontarget species are listed in Table 3.6. Additional toxicity data on neem products are found in several reviews (Kreutzweiser, 1997; Raguraman et al., 2004; Stark, 2006).

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Table 3.6  Acute Toxicity of Neem Preparations Test material

Species, route

LD50 (g/kg)

Neemrich I

Rat, oral

8.7

Sharma et al. (1984)

Rat, oral

a

Gandhi et al. (1988)

Neem oil b

Other data

14

a

References

Rat, oral

5

Larson (1989)

Rat, oral

10

Trifolio-M (1995)

Rat, dermal

11.2

Sharma et al. (1984)

Rat, oral

1.57

Mahboob et al. (1998)

Leaf and bark extract

Mouse, oral

13

Okpanyi and Ezeukwu (1981)

Neemrich I

Mouse, oral

6.8

Sharma et al. (1984)

Mouse, oral

10

Trifolio-M (1995)

Margosan-O

NeemAzal-T/S

c

Neemrich I Vepacide

d

NeemAzal-T/S

c

Neem oil

Rabbit, oral

24

a

Gandhi et al. (1988)

Margosan-O

b

Rabbit, dermal

LC50  2 ml/kg

Larson (1989)

Margosan-O

b

Rabbit, 1-h inhalation test

LC50  43.9 mg/l

Larson (1989)

Neemrich I

Chicken, oral

39.9

Sharma et al. (1984)

Neemrich I

Pigeon, oral

6.3

Sharma et al. (1984)

Margosan-Ob

Bobwhite quail, 5-day

LC50  7000 ppm

Larson (1989)

Margosan-Ob

Mallard duck, 5-day

LC50  7000 ppm

Larson (1989)

Margosan-Ob

Rainbow trout, 96-h

LC50  8.8 ppm

Larson (1989)

Margosan-Ob

Rainbow trout, 96-h

LC50  29 ppm

Wan et al. (1996)

Azatine

Rainbow trout, 96-h

LC50  4 ppm

Wan et al. (1996)

Azadirachtin

f

Rainbow trout, 96-h

LC50  4 ppm

Wan et al. (1996)

Azadirachtin

g

Rainbow trout, 96-h

LC50  61 ppm

Wan et al. (1996)

b

Coho salmon, 96-h

LC50  38 ppm

Wan et al. (1996)

Coho salmon, 96-h

LC50  5 ppm

Wan et al. (1996)

Coho salmon, 96-h

LC50  81 ppm

Wan et al. (1996)

Aqueous leaf extract

Prochilodus lineatus, 24-h

LC50  4800 ppm

Winkaler et al. (2007)

Margosan-Ob

Bluegill, 96-h

LC50  37 ppm

Larson 1989

Azadirachtin (95% pure)

Crayfish, 96-h

LC50  1 ppm

Goktepe and Plhak (2004)

Bioneemh

Margosan-O Azatin

e

Azadirachtin

g

Crayfish, 96-h

LC50  5180 ppm

Goktepe and Plhak (2004)

Neemix

i

Grass shrimp, 96-h

LC50  1520 ppm

Goktepe and Plhak (2004)

Neemix

i

Blue crab, 96-h

LC50  460 ppm

Goktepe and Plhak (2004)

Neem stem bark extract

Aphyosemon giardneri, 96-h

LC50  15.1 ppm

Osuala and Okwuosa (1993)

Neem stem bark extract

Bulinus truncatus, 96-h

LC50  11 ppm

Osuala and Okwuosa (1993)

Margosan-Ob

Daphnia magna, 48-h

LC50  13 ppm

Larson (1989)

Daphnia magna, 48-h, pH  7–8

EC50  125 ppm

Scott and Kaushik (1998)

Margosan-O

b

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Table 3.6  (Continued) Test material

Species, route

Margosan-Ob

Other data

References

Daphnia magna, 48-h, pH  8–9

EC50  923 ppm

Scott and Kaushik (1998)

Azadirachtin (95% pure)

Daphnia pulex, 48-h

LC50  0.382 ppm

Goktepe and Plhak (2002)

Bioneemh

Daphnia pulex, 48-h

LC50  33 ppm

Goktepe and Plhak (2002)

Neemixi

Daphnia pulex, 48-h

LC50  28 ppm

Goktepe and Plhak (2002)

Neemixj

Daphnia pulex, 48-h

LC50  0.68 ppm

Stark and Banks (2001)

LC50  1.12 ppm

Kreutzweiser (1997)

Azatin

k

Mayfly, 1-h

Neemrich I Neem oils

m

LD50 (g/kg)

l

Honeybee, topical

0.0735

Honeybee, 72-h contact

Sharma et al. (1984) 2

LD30  0.11–0.53 mg/cm

Melathopoulos et al. (2000)

a

Milliliters per kilogram. Formulation containing 14–20% neem oil giving a total of 3 g/l azadirachtin. Formulation containing 1% azadirachtin. d Contains 12% azadirachtin and 88% other neem constituents. e Formulation containing 3% azadirachtin and approximately 27% other neem constituents. f Contains 49% azadirachtin and 51% other neem constituents, including salannin, nimbin, etc. g Formulation containing 4.6% azadirachtin and 15% other neem constituents. h Formulation containing 0.09% azadirachtin. i Formulation containing 0.25% azadirachtin. j Formulation containing 4.5% azadirachtin. k Contains 3% azadirachtin. l Exposure for 1 h in a flow-through test, followed by a 21-day period for mortality observation. m Various commercial neem oils with unspecified azadirachtin-content were applied to the bottom surface (28.3 cm2) of the Petri dish cage. b c

Neem seed-based animal feed supplements were found to be safe to chicken and cattle but not to sheep. Neem leaf was reported to be toxic to sheep (Ali and Salih, 1982), to goats and guinea pigs (Ali, 1987) but not to rabbits (Thompson and Anderson, 1978). Based on a threegeneration toxicology study with rats, debitterized neem oil, which is obviously depleted of limonoid ingredients, was recommended as suitable for human consumption (Chinnasamy et al., 1993). Methanolic extracts of neem leaf and bark had oral LD50 values of about 13  mg/kg in mice, and the poisoning signs were discomfort, gastrointestinal spasms, loss of appetite, hypothermia, and, ultimately, convulsion leading to death within 24 h (Okpanyi and Ezeukwu, 1981). The toxicity of Vepacide, an enriched neem oil-based preparation containing 12% azadirachtin plus additional terpenoids, upon oral administration of 80, 160, and 320 mg/kg daily doses for 90 days, was studied in male rats (Mahboob et al., 1998). On the 90th day, the high and medium doses caused significant decreases in: (1) CYP concentration in the liver, lungs, and kidneys but not the brain; (2) cytochrome b5 in the brain; and (3) CYP reductase level in the liver and brain. The highest dose also caused 10% mortality; the medium dose elicited toxic signs, including behavioral abnormalities, lacrimation, reduced feeding, and loss in body weight, but no toxicity was seen for the low dose. The toxic symptoms disappeared

upon cessation of treatment. By contrast, a similar 90-day study by Raizada et al. (2001) using a neem extract containing 12% azadirachtin did not find any histopathological, hematological, enzymatic or other adverse effects at 500, 1000 and 1500 mg/kg daily oral doses; a single oral dose of 5000 mg/kg of this substance produced neither toxic symptoms nor death. Irritation  A commercial azadirachtin-enriched pulverized neem extract caused only slight and transient erythema in an acute dermal irritation study in rabbits; however, the pronounced skin irritation observed in the guinea pig indicates that this technical product is a dermal sensitizer (U.S. EPA, 2008b). Cytotoxic effects  In insect and mammalian cell culture tests, nimbolide was found to be the most active ingredient among the cytotoxic limonoids of neem seed extracts. Azadirachtin as well as neem oil containing less than 1% of total limonoids were essentially devoid of cytotoxicity in this study (Cohen et al., 1996a, 1996b; see also Cui et al., 1998; Kigodi et al., 1989). Azadiracthin, however, was cytotoxic to several human glioblastoma cell lines by reducing cell survival and preventing mitosis (Akudugu et al., 2001). A recent work has found that azadirachtin disrupts the distribution of mitotic spindles and cytoplasmic microtubules during gametogenesis of the malarial parasite Plasmodium berghei (Billker et al., 2002).

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Daily oral doses of 0.5–2.0 g/kg of an ethanolic neem leaf extract of unknown composition to male mice showed cytotoxicity and caused chromosomal abnormalities in spermatocytes after 7 days of treatment (Khan and Awasthy, 2003). On the other hand, neem leaf preparations have repeatedly shown cancer chemopreventive action both in vitro and in vivo (see, for example, Haque and Baral, 2006; Manikandan et al., 2008). In addition, neem leaf extract at 1–3 ppm ameliorated the chromosome aberrations and micronucleus formation in fish exposed to 0.6 ppm pentachlorophenol or 75 ppm 2,4-D under laboratory conditions (Farah et al., 2006). (e)  Toxic Effects in Humans Poisoning incidents  Neem seed oil produced occasional diarrhea, nausea, and general discomfort when given orally as an anthelmintic (see Jacobson, 1989). Neem leaf extract-poisonings with ventricular fibrillation and cardiac arrest have occasionally been reported (Balakrishnan et al., 1986). Sinniah and Baskaran (1981) summarized 13, including two fatal, poisoning cases due to neem seed oil. Five to ten milliliters of this traditional remedy, also called margosa or kohomba oil, given orally to children against minor ailments caused vomiting, drowsiness, tachypnea with acidotic respiration, polymorphonuclear leukocytosis, and encephalopathy that developed within hours of ingestion. Seizures, associated with coma, also developed in some cases. Autopsy demonstrated pronounced fatty acid infiltration of the liver and proximal renal tubules, with mitochondrial damage and cerebral edema, changes consistent with Reye’s syndrome. Follow-up model studies with mice suggested that the syndrome could be associated with long-chain fatty acid or lipid components of the oil provoking mitoses of hepatocytes within 30 min after ingestion, hypertrophy of endoplasmic reticulum, and loss of liver glycogen, which was consistent with fat accumulation in the liver cells (Sinniah et al., 1989). In model studies with rat liver, Trost and Lemasters (1996) proposed that the pathogenesis of Reye’s syndrome, caused by chemicals such as salicylic acid, valproic acid and neem oil, is associated with the induction of mitochondrial permeability translation. The induction effect could be blocked by cyclosporin A. Dhongade et al. (2008) have described a severe poisoning case of a 5-year-old boy who presented refractory epileptic seizures, dilated pupils, tachycardia, and dyspnea 1 h after accidental ingestion of neem oil; metabolic acidosis was also found. He required resuscitation and mechanical ventilation as well as intravenous diazepam to control convulsions. His general conditions improved over several days, but neurodeficits and choreoathetoid movements persisted for months. A similar child poisoning without sequelae was described by Sri Ranganathan et al. (2005).

Although neem preparations have been used safely against a variety of skin diseases, occasional dermatitis in sensitive individuals have been reported (Reutemann and Ehrlich, 2008). Treatment  The recommended treatment for neem oil poisoning is control of seizures by diazepam, respiratory support, correction of acidosis, reduction of cerebral edema by dexamethasol and/or mannitol, and hydration of the patient (Sinniah and Baskaran, 1981).

3.2.2  Microbial Insecticides 3.2.2.1  Bacillus thuringiensis Endotoxins (a)  Introduction Bacillus thuringiensis (Bt) is an aerobic, sporeforming, Gram-positive, rod-shaped bacterium distributed widely in the natural environment from the Arctic to the Tropics (Martin and Travers, 1989). The entomopathogenic and insecticidal action (“sotto” disease) of the bacterium was first noted by Ishiwata in Japan in 1901. In 1915, another strain was found in Thuringia, Germany, by Berliner, who named it B. thuringiensis. The first Bt-based microbial insecticide was commercialized in France in 1938; Bt was registered in the United States in 1961. Thorough overviews on the history, biology, early and current use, and (eco)toxicology of this most important microbial insect control agent are available (Bravo et al., 2005; Burges, 1981; Glare and O’Callaghan, 2000). This section will address issues related to the conventional use of this bacterial insecticide and genetically engineered crop plants expressing Bt insecticidal toxins will only be mentioned when appropriate. (b)  Identity, Properties, and Nomenclature During the past century, thousands of Bt bacterium isolates have been obtained from sources as diverse as living and dead insects, soil, plants, grain dust or flour, and water (Bernhard et al., 1997; Chaufaux et al., 1997). The bacterium was isolated also from rodents and insectivorous small animals living in the wild (Swiecicka et al., 2002) and from decaying leaf litter in mosquito breeding areas (Tilquin et al., 2008). The Bt isolates can be divided into at least 80 serologically different Bt subspecies (or varieties), each producing distinct crystalline inclusions during sporulation. The crystals (or -endotoxins) are selectively toxic against many insect pests and disease vectors. The following Bt subspecies have practical importance: kurstaki and aizawai (against lepidopteran larvae), israelensis (against mosquitoes and blackflies), and morrisoni (or tenebrionis) and japonensis (against beetles). Each Bt subspecies may synthesize more than one class of -endotoxin. The insect-specific Bt toxins were formerly classified according to their host, size, and crystal shape into four

Chapter | 3  Pest Control Agents from Natural Products

major groups (Höfte and Whiteley, 1989): CryI (targeting Lepidoptera), CryII (targeting both Lepidoptera and Diptera), CryIII (targeting Coleoptera), and CryIV (targeting Diptera). The recently proposed nomenclature uses Arabic numerals (e.g. Cry1), is based solely on amino acid identity, and allows phylogenetic clustering currently into 40 systematically arranged primary ranks (Crickmore et al., 1998, 2009). In addition to Cry toxins, some Bt contain another endotoxin (Cyt), which has a specific action on Diptera in vivo and a broad activity spectrum in vitro, being cytolytic and/or hemolytic for most eukaryotic cells, including horse, sheep, rat, mouse, rabbit, and human erythrocytes (Guerchicoff et al., 2001; Knowles et al., 1992; Thomas and Ellar, 1983). It appears that Cry and Cyt proteins interact synergistically in insect gut to exert their biological effects. Additional, structurally unrelated proteins produced by the bacterium are the vegetative insecticidal proteins (Vip) (Espinasse et al., 2003; Estruch et al., 1996), the antimicrobial bacteriocins (see, for example, Jung et al., 2008), the antifungal kurstakin lipopeptides (Hathout et al., 2000), and the noninsecticidal but cytotoxic parasporins (Mizuki et al., 2000). Among these, the Vip exotoxins produced during the vegetative phase of growth are of practical importance. Along with -endotoxins, many variants of Bt, subsp. aizawai in particular, produce a low-molecular-weight, thermostable, and water-soluble insecticidal compound, known as -exotoxin [23526-02-5] (see Perani et al., 1998; Šebesta et al., 1981). This exotoxin, also called thuringiensin, is a structural analog of ATP (Espinasse et al., 2003; Farkaš et al., 1969) (Figure 3.7) and inhibits DNA-dependent RNA polymerases both in prokaryotic and eukaryotic cells (Bond et al., 1969; McClintock et al., 1995; Šebesta and Horská, 1970). Although -exotoxin-based insecticides have been commercialized to control flies (see Abrosimova et al., 1985; Haufler and Kunz, 1985; Hernández et al., 2001; Hsu et al., 1997), the endotoxin-based Bt products are required to be free from the nucleoside analog. Bt is taxonomically closely related to the mammalian pathogens Bacillus cereus and Bacillus anthracis (VilasBôas et al., 2007), which, however, lack the -endotoxin and are thus noninsecticidal and can be harbored in insects. However, the -endotoxin crystal-coding Bt plasmid could be transferred to B. cereus, yielding transcipients that produced crystals of the same antigenicity as the donor strain (González et al., 1982). Rapid and highly sensitive analytical methods, based either on mass spectroscopy (Ullom et al., 2001), polymerase chain reaction (see, for example, Henderson et al., 1994; Song et al., 2006), optical chromatography (Hart et al., 2006), or immune reaction (Campbell and Mutharasan, 2007), have been developed to distinguish between Bt and the biological warfare agent B. anthracis.

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(c)  Structure Comparative analyses of the structure of Bt toxins have provided insight into the mode of action and specificity of these proteins. The molecular weights of -endotoxins vary from 27 to 140 kDa, with regions of amino acid sequence homology or high similarity interspersed with variable regions (Höfte and Whiteley, 1989; Pigott and Ellar, 2007). The short segment at the C-terminal of the protein is not required for toxicity and is thought to play a role in crystal formation within the bacterium. The X-ray crystal structures of the following toxins have been reported: the Coleoptera-specific Cry3Aa (Li et al., 1991) and Cry3Bb1 (Galitsky et al., 2001), the Lepidopteraspecific Cry1Aa (Grochulski et al., 1995) and Cry1Ac (Li et al., 2001), the Lepidoptera/Diptera-specific Cry2Aa (Morse et al., 2001), and the Diptera-specific Cyt2Aa (Li et al., 1996; Morse et al., 2001), Cry4Aa (Boonserm et al., 2006) and Cry4Ba (Boonserm et al., 2005) have been solved. The known Cry structures show a common topology composed of three domains. Domain I at the N-terminal forms a bundle of -helices and is responsible for pore formation. Domains II and III consist of several -sheets that form a -prism and a -sandwich, respectively, and both domains are involved in receptor binding. For a more detailed analysis, see the review by Pigott and Ellar (2007). All the insecticidal endotoxins of Bt appear to be encoded by a single megaplasmid (pBtoxis), the complete nucleotide sequence of which has been determined (Berry et al., 2002). (d)  Formulations and Uses Bt insecticides are produced by fermentation and the products usually consist of cells, spores, and parasporal crystals. Commercial preparations contain at least 1012 spores/l but because spore counts do not always correlate with the quantity of -endotoxin, thus with insecticidal activity, standardization of the formulations by bioassay is necessary. This is done by comparing the LC50 of the actual sample with the LC50 value of a standard against the target species and the potency of the sample is then expressed in international unit (IU) per milligram or milliliter COOH H

OH

H

OH O

H

O

H

O

NH2

OH N

P(OH)2 O O

O

N N

COOH OH

OH OH

OH

β-exotoxin Figure 3.7  Structure of Bacillus thuringiensis -exotoxin.

N

156

Hayes’ Handbook of Pesticide Toxicology

of product. The toxin concentration depends on the formulation and typically ranges from several thousands to hundred thousand IU/mg. The preparations are formulated as water-soluble, floating or frozen granules, liquid concentrates, emulsifiable suspensions, wettable powders, and slow release rings or tablets. Repeated applications at rates ranging from 0.1 to 7 kg/ha of formulated product are used. Due to the narrow larvicidal activity spectrum of the endotoxins, any given preparation can only be used against a few selected lepidopteran, dipteran, or coleopteran species. Bt preparations are used to control various insect pests in cotton, vegetables, orchards, maize, forests, turf, and ornamentals, as well as against mosquito and blackfly larvae in water bodies and sewage filters. Suspension formulations containing Lepidoptera-specific Cry1 toxins encapsulated in killed Pseudomonas fluorescens are also available (Copping, 2004). Several modified cry genes that code for Cry proteins, often in a truncated form, have been engineered, along with marker proteins, gene expression promoters, terminators and enhancers into crops, including maize, cotton, eggplant, rice, soybean and potato varieties, and such crops have been available commercially since 1996. Variants of vip3A genes encoding insecticidal Vip3A proteins have also been inserted into genomes of cotton and maize. Recently, two or more genes coding different toxins have been engineered together into plants (gene stacking/pyramiding). For example, transgenic maize that contains plasmids with the cry1Ab and cry3Bb1 genes coding for Cry1Ab and Cry3Bb1 proteins, respectively, is endowed with protection against both corn borers (Lepidoptera) and corn rootworms (Coleoptera). Stacking different types of genes into maize confers combined herbicide and insect tolerance. The complete toxicological evaluation of the transgenic plants continuously producing the proteinaceous insecticide at varying (1–100 ppm) levels in different plant tissues, however, is a challenging task requiring the development of test methodologies and safety criteria different from those used for the safety assessment of conventional chemical insecticides (Mendelsohn et al., 2003; Organisation for Economic Co-operation and Development, 2007; Romeis et al., 2006). By combining genetic elements that encode Cyt and Cry toxins from Bt subsp. israelensis and the binary endotoxin from B. sphaericus, itself a mosquito larvicide (Lacey, 2007), Federici et al. (2007) have recently succeeded in creating recombinant bacterial strains with mosquitocidal activity tenfold higher than the conventional strains of these bacteria.

its toxin can be inactivated by usual physical (for example, heating at 60 to 90°C) and chemical (formaldehyde, chlorine, chlorine dioxide, hypochlorite, or strong acid solution) sterilization methods. Formulated Bt preparations can be stored for at least a year when kept at 20–25°C but freezing of liquid formulations for several months results in substantial loss of activity (see, for example, Boisvert and Boisvert, 2001). Under field conditions, Bt spores and crystals have, in general, low persistence, depending on the type of formulation, on the rate of degradation by microorganisms and the target insects, and on abiotic factors. Field half-lives ranging from 0.5 to 4 days were reported for various preparations (Aldemír, 2007; Beegle et al., 1981; de Lara Haddad et al., 2005; Herman et al., 2002; Ignoffo et al., 1974, 1977). Sunlight alone or in combination with high temperature and rain is responsible for the rapid inactivation of Bt spores in the field (Ishiguro and Miyazono, 1982; Leong et al., 1980; Raun et al., 1966; van Frankenhuyzen and Nystrom, 1989). Spectroscopic studies with purified protein crystals pointed out the role of exogenous photosensitizers, such as singlet oxygen, in the breakdown of tryptophan side chains in light-initiated inactivation (Pozsgay et al., 1987). In the soil, under certain conditions the toxin could persist for several months or longer by binding on clays and humic acids and this, on the one hand, improves its insect controlling effect but, on the other, could enhance the development of resistance in the target species and may present a potential hazard to nontarget organisms (reviewed by Stotzky, 2004). For example, Bt spores were found to persist in forest soil for up to 2 years after a 5-year intensive use of the insecticide against the gypsy moth (Smith and Barry, 1998). No -endotoxin could be detected in crops newly planted in toxin-containing soil indicating the lack of systemicity of the protein (see Stotzky, 2004). Efficacy studies with Bt subsp. israelensis in aquatic environments have shown that the larvicidal activity of most preparations lasts only for a week unless specially formulated (see Kahindi et al., 2008 and references therein). The persistence of the insecticidal crystals in such environments is less studied. Using a suspension formulation of Bt subsp. israelensis in an experiment carried out in a low-temperature pond, Boisvert and Boisvert (1999) followed changes of insecticidal activity and the dynamics of crystal deposition. The toxicity of the liquid fraction decreased over time with periphyton, sediment and vegetation apparently acting as sink. The toxin crystals adsorbed onto the vegetation remained active for up to 22 weeks but recycling of the toxin from killed larvae was negligible.

(e)  Stability The proteinaceous Bt -endotoxin is not soluble in water and organic solvents but can be dissolved in dilute alkalis such as aqueous NaOH solutions. However, Bt-containing products lose activity at pH  8. Both the Bt bacterium and

(f)  Biological Properties Mode of action  The insecticidal mode of action of Bt spores and their toxins is complex and still controversial (reviewed by Bravo et al., 2005, 2007; Pigott and Ellar, 2007; Schnepf et al., 1998). Bt-based insecticides kill

Chapter | 3  Pest Control Agents from Natural Products

insects not because of their infectivity but because of their Cry and Cyt endotoxins that disrupt cell membranes in the midgut epithelium of susceptible insects. (It appears though that other protein components of the spore appear to contribute to the insecticidal effect.) It is generally assumed that portions of the endotoxins form transmembrane pores in the target cells. For Cry proteins, the -helix domains form a pore, while for Cyt toxins it is the -barrel regions that are involved in pore formation. The pores then allow ions and water to leak into the cells, resulting in swelling, lysis, and the eventual death of the insect. According to the currently prevailing model proposed for the molecular mechanism of action of Cry toxins, Cry1A in particular, the four critical sequential steps involved are as follows. 1. Ingestion. Cry toxins have no contact activity, they must be ingested to be toxic. 2. Solubilization and proteolytic activation. Within the digestive tract, the crystals dissolve in the alkaline (lepidopteran and dipteran gut) or the neutral/acidic (coleopteran gut) environment. The bacterium-produced long protoxin is then activated in the midgut lumen by proteolytic removal of a short peptide fragment typically at the N-terminal. (Accordingly, to exhibit toxicity in vitro, the protoxin requires prior enzymatic activation.) 3. Binding to target site(s). The activated 60–70 kDa toxin then selectively and reversibly binds to insect-specific receptor(s) in the microvilli of the midgut epithelial cells. Many putative Cry toxin receptors have been reported, of which the best characterized are the aminopeptidase N receptors, the cadherin-like receptors, an alkaline phosphatase receptor, a 270-kDa glycoconjugate, and a 252-kDa protein. In nematodes, certain glycolipids seem to function as Cry receptor. 4. Formation of toxic lesions. After binding to membrane receptor(s), the activated toxins undergo conformational changes then presumably oligomerize before rapid and irreversible insertion into the membrane to form a pore. The pore increases cell membrane permeability and initiates an influx of cations, especially K. Osmotic balance is thus disturbed, which causes the cell to swell and burst by a process called colloid-osmotic lysis, eventually leading to larval death within 1 or 2 days. With accumulating new information on the molecular aspects of Cry action, additional and/or more refined models for the insecticidal action of the toxins are being developed (for a discussion of these, see Pigott and Ellar, 2007; Soberón et al., 2009). For example, Broderick et al. (2006) have recently provided evidence that in gypsy moth enteric bacteria are responsible for the septicemia associated with the insecticidal effect of ingested Bt. The Vip proteins disrupt insect midgut epithelial cell wall integrity similarly to Cry toxins but involve a different molecular target (Lee et al., 2003a).

157

Distribution and excretion  The toxicity, fate, and infectivity of Bt subsp. israelensis preparations in mice, rats, and rabbits have been studied in detail by Siegel et al. (1987). Viable bacteria could be recovered at the injection site and from the spleen of mice 14 days after subcutaneous administration of 109 Bt organisms. Aerosol exposure of rats to a spray containing 2.05  106 bacterial organisms/ml for 30 min showed that viable Bt from the lungs cleared completely within 7 days without any lesions, and no bacteria could be detected in the spleen. Twentysix of 42 athymic mice died within 5–10 h after receiving 3.4  107 bacteria intraperitoneally; colony-forming units (CFUs) in the spleen of the surviving animals declined with time but persisted as long as 7 weeks. No mortality was seen in euthymic mice that received a comparative dose. In rats, intracerebral injection experiments with different Bt preparations only the highest dose of 107 Bt organisms per animal produced mortality (79–83%), and clearance from the spleen and brain of animals receiving 1.15  105 bacteria was essentially complete within 3 weeks. Because recovery of CFUs decreased rather than increased over time in all experiments, the Bt subsp. israelensis preparations tested were clearly not infective. The involvement of phagocytic cells such as macrophages, the lymphatic system, and the blood stream in clearance of the bacteria was also proposed. Subsequent tests with mice and rabbits confirmed and extended these findings (Siegel and Shadduck, 1990; see also Siegel, 2001). On intraperitoneal injection into mice, Bt subsp. israelensis CFUs were also recovered from heart blood, and their disappearance from it coincided with their clearance from the spleen. Immunodeficient mice cleared Bt preparations at a slower rate. However, mice failed to remove one preparation of Bt subsp. israelensis from their enlarged spleen and a constant number of bacteria (1.6–20  106 CFUs) was recovered even after 10 weeks. On ocular installation, the bacteria persisted in both flushed and unflushed rabbit eyes for 1 week only. In a series of experiments with various Bt preparations (Bishop et al., 1999), oral daily dosages of 5  1010 Bt spores, containing endotoxins, -exotoxin as well as cytolytic enterotoxins, to rats for 3 weeks resulted in no toxic symptoms. Furthermore, in agreement with previous findings, there was no evidence of bacterial infection indicating that the spores are uncapable of germination in the gut of this rodent. Subcutaneous injection of a single dose of 5  1010 spores showed no ill effects either and no bacteria could be recovered from the blood of animals 3 weeks after treatment. From spinach leaves sprayed with a Bt subsp. kurstaki preparation, only brief treatment in boiling water removed (dissolved) spores capable of colony forming; simple washing with cold running water reduced only about 50% of the spore load. McClintock et al. (1995) (see also U.S. EPA, 1998b) reviewed unpublished studies on the clearance of viable

158

Bt spores from rodents. Microbial clearance through the digestive tract of rats was complete in some instances in 2 days. Clearance of inhaled Bt subspecies aizawai and kurstaki from mouse brain, blood, liver, kidney, lung, lymph nodes, and spleen was complete in 2–3 weeks. Intravenous doses cleared at a slower rate from these tissues in mice and a similar pattern was observed for rats; viable CFUs could be recovered even after 50 days in some of the tissues of these rodents. In a recent interesting study by Wilcks et al. (2006), rats were first nourished to have a digestive tract simulating human microflora then received daily oral treatment of 107–108 spores or vegetative cells of commercial Bt subsp. israelensis or subsp. kurstaki preparations. Analysis of fecal samples of the animals after 4 days of treatment showed, in general, that only the spores but not the vegetative cells survived the gastric passage. In one treated animal, however, spores of subsp. kurstaki could germinate in the small intestine and resporulate in the large bowel before excretion with the feces. Translocation to liver and spleen was also observed in this animal, indicating that Bt is capable of crossing the intestinal barrier in certain cases. None of the animals showed adverse effects or substantial changes in their gut microflora. Studying the uptake and excretion of Bt subsp. israelensis spores in the fathead minnow, Snarski (1990) reported that upon 1 h exposure of fish to 2.2  105 CFU/ml, a spore density hundredfold higher than the recommended application rate, a whole-body count of 4.4  106 CFU per fish was found. Analysis of whole-body, gill and gastrointestinal tract homogenates indicated that mostly ingestion accounted for spore accumulation. Two days after treatment the whole-body spore count decreased to an average of 3 CFU per fish while fecal excreta continued to contain high though gradually decreasing levels of spores throughout the first week of postexposure (from 105 CFU per fish to 103 CFU per fish). The fish showed no signs of toxicity at 6.4  105 CFU/ml or lower application rates. (g)  Toxicity to Animals and Laboratory Studies Acute toxicity of endotoxins  As mentioned before, the crystalline Bt endotoxins require activation by alkalis and/ or digestion, conditions absent in the mammalian stomach but present in the insect midgut, providing a basis for selective toxicity. For example, in mice, intravenous administration of crystalline -endotoxin of Bt subsp. israelensis at 1 mg per animal produced no toxic symptoms, but the solubilized toxin had an LD50 of approximately 0.49 mg per animal (Thomas and Ellar, 1983). Representative acute toxicity data are given in Table 3.7 and by Lamanna and Jones (1963). The mammalian toxicology studies submitted to the EPA on Bt-based insecticides were summarized by McClintock et al. (1995) and relate the now internationally used CFU dose data to exposure doses used in earlier

Hayes’ Handbook of Pesticide Toxicology

studies. A CFU is defined as a single, viable propagule that produces a single colony (a population of the cells visible to the naked eye) on an appropriate semisolid growth medium. Due to their narrow spectrum of insecticidal activity, Bt preparations have been used safely and reliably, even in delicate forest and aquatic ecosystems. The effects of various Bt preparations on nontarget aquatic and terrestrial organisms were recently compiled (Boisvert and Boisvert, 2000; Glare and O’Callaghan, 2000; see also Jackson et al., 2002; U.S. EPA, 1998b; World Health Organization, 1999). Bt preparations containing 0.1% of the water-soluble -exotoxin had an intraperitoneal LD50 value of 364– 387 mg/kg (5.9–6.2  109 cells/kg) in rats but were devoid of acute toxicity at 10,000 mg/kg (1.6  1011 cells/kg) oral, at 6000 mg/kg (9.6  1010 cells/kg) dermal, and at 300 mg/ m3 (4.8  1010 cells/kg) inhalation doses (Khalkova et al., 1993). Chronic toxicity  In rat chronic studies with Bt subsp. kurstaki, daily doses of 8.4 g/kg by oral administration for 90 days and feeding for 2 years did not show treatment-related effects. A 13-week study with Bt subsp. kurstaki (Dipel), daily oral administration of 1.3  109 spores/kg, showed no toxicity or infectivity in rats. With Bt subsp. israelensis, the no-observed-effect level (NOEL) was a daily dose of 4 g/kg in a 3-month study with rats (McClintock et al., 1995; see also California Department of Food and Agriculture, 1998). Sheep fed with 500 mg/kg daily doses of Bt subsp. kurstaki insecticides (Dipel or Thuricide) for 5 months showed no treatment-related effects (Hadley et al., 1987), although the bacterium could be cultured from blood and tissue samples taken at the end of the trial. The only pathological finding was mild lymphocytic hyperplasia in Pleyer’s patches of the cecum of some animals. Irritation and sensitization  No allergenic response to Thuricide Bt preparation was evident in mice by inhalational exposure to 9  1010 viable spores for 10 min, and in guinea pigs by subcutaneous injection of 10 doses of approximately 9  105 spores during 3 weeks or topical application of approximately 4.5  107 spores to intact or abraded skin (Fisher and Rosner, 1959). Bt subsp. israelensis dry-powder preparations caused slight ocular irritation, whereas pastes caused severe conjunctival congestion and corneal injury in rabbit eyes (Siegel and Shadduck, 1990; Siegel et al., 1987). Based on animal experiments and in vitro digestibility tests using conditions that simulate gastric fluid as well as on amino acid sequence comparisons with known allergens, it is assumed that Bt proteins behave as dietary proteins and have no allergenic potential in humans when ingested (see Bernstein et al., 2003; Mendelsohn et al., 2003).

Chapter | 3  Pest Control Agents from Natural Products

159

Table 3.7  Acute Toxicity of Bacillus Thuringiensis Preparations Bt subspecies, toxin type

Test species, route

LD50

israelensis

Rat, oral

2670 mg/kg

McClintock et al. (1995)

israelensis

Rat, oral

1.2  1011 spores/kg

McClintock et al. (1995)

israelensis

Rat, dermal

2000 mg/kg

McClintock et al. (1995)

israelensis -endotoxina

Rat, ip

1.95 mg/kg

Roe et al. (1991)

israelensis -endotoxina

Rat, iv

21 mg/kgb

Roe et al. (1991)

israelensis -endotoxin

a

Rat, sc

Other data, comments

References

b

Roe et al. (1991)

9 mg/kg

7

israelensis

Rat, inhalation

israelensis -endotoxina

Mouse, male, oral

30 ppm

Cheung et al. (1985)

israelensis -endotoxina

Mouse, male, ip

1.31 ppm

Cheung et al. (1985)

israelensis -endotoxinc

Mouse, male, ip

0.77 ppm

Mayes et al. (1989)

israelensis -endotoxina

Mouse, male, ip

2.33 mg/kg

Mayes et al. (1989)

israelensis -endotoxina

Mouse, iv

~16 mg/kgd

Thomas and Ellar (1983)

israelensis -endotoxin

e

israelensis -endotoxin

a

israelensis israelensis israelensis -endotoxin

a

israelensis -endotoxin

a

Mouse, iv Suckling mouse, sc

LC50  8  10 spores/rat

McClintock et al. (1995)

b

Thomas and Ellar (1983)

33 mg/kg

f

2.7–6.6 mg/kg

Thomas and Ellar (1983)

9

Rabbit, oral

2  10 spores/rabbit

McClintock et al. (1995)

Rabbit, dermal

6280 mg/kg

McClintock et al. (1995)

Japanese quail, ip

22.8 mg/kg

Kallapur et al. (1992)

b

Japanese quail, intranasal

50 mg/kg

Kallapur et al. (1992)

israelensis -endotoxina

Quail, iv

100 mg/kgb

Roe et al. (1991)

israelensis -endotoxina

Quail, sc

100 mg/kgb

Kallapur et al. (1992) 6

israelensis

Fathead minnow, 24-h

LC100  6.5  10   CFU/mlg

Snarski (1990)

israelensis

Brook trout, 48-h

LC50  2321 ppm

Wipfli et al. (1994)

israelensis

Brown trout, 48-h

LC50  1691 ppm

Wipfli et al. (1994)

kurstaki

Rat, oral

4.7  1011 spores/kg

McClintock et al. (1995)

kurstaki

Rat, dermal

3.4  1011   spores/kg

McClintock et al. (1995)

kurstaki -endotoxina

Mouse, iv

33 mg/kgb

Thomas and Ellar (1983)

kurstaki -endotoxine

Mouse, iv

33 mg/kgb

Thomas and Ellar (1983)

kurstaki -endotoxina

Mouse, male, oral

30 ppm

Cheung et al. (1985)

Mouse, male, ip

30 ppm

Cheung et al. (1985)

kurstaki -endotoxin

a

kurstaki

Rat

kurstaki

Daphnia, 21-day

kurstaki, Cry1Aa

Zebrafish, 96-h

aizawai

Rat

aizawai

Daphnia, 21-day

8

McClintock et al. (1995)

10  CFU/animal

8

LC50  5–50 ppm

U.S. EPA (1998b)

LC50  85.9 ppm

Grisolia et al. (2009)

b

McClintock et al. (1995)

2  10   CFU/animal

EC50  0.8–2.7 ppm

U.S. EPA (1998b) (Continued)

Hayes’ Handbook of Pesticide Toxicology

160

Table 3.7  (Continued) Bt subspecies, toxin type

Test species, route

LD50

tenebrionis

Rat

2  108  CFU/animal

tenebrionis

Daphnia, 48-h

h

Other data, comments

References McClintock et al. (1995)

EC50  100 ppm 8

U.S. EPA (1998b)

Rat, oral

LD50  1.2  10   spore/ animal

Meher et al. (2002)

kenyaeh

Rabbit, dermal

LD50  2.5  107  spore/ animal

Meher et al. (2002)

kenyaeh

Mosquitofish, 96-h

BMB696B Vip3A toxini

Rat, oral

kenyae

Vip3Aa protein

Mouse, oral

LC50  2.5  1010  spore/l

Meher et al. (2002)

LD50  5000 mg/kgb

Peng et al. (2007)

b

U.S. EPA (2008c)

LD50  3675 mg/kg

a

Solubilized endotoxin. No mortality. A 28-kDa polypeptide fraction of solubilized endotoxin. d Calculated from original mortality data. e Crystalline endotoxin. f Lethal doses. g Mortality was attributed to the indirect effect of dissolved oxygen depletion due to formulation components. h Suspension of a wettable powder formulation. i Genetically modified Bt strain preparation containing Vip3Aa7 protein was used. b c

Genotoxicity, mutagenicity, and cytotoxicity  At 0.5–1000 g/plate application rates, the endotoxin-based Thuricide was not mutagenic in the 48-h S. thyphimurium mutagenicity assay with or without a metabolic activator, although the validity of the Ames test for this type of product was questioned (California Department of Food and Agriculture, 1998). The Ames test showed no mutagenicity for -exotoxin (Carlberg et al., 1995). Ren et al. (2002) examined the genetic toxicity of injected Bt subsp. kurstaki preparation to the fifth instar nymph of the grasshopper Oxya chinensis. At one-tenth of the LD50 dose (5.50 IU/ml) chromosome and chromatid breaks were observed, although their incidence was relatively low compared to the effect of cyclophosphamide. However, the genotoxic effect of oral treatment, which is more relevant to the mode of toxic action of Bt endotoxin, was not examined in this study. Grisolia et al. (2009) used the zebrafish model for the evaluation of the acute and genotoxic effects of Cry1Aa, Cry1Ab, Cry1Ac, and Cry2A toxins alone and as binary mixtures. In adult fish exposed to 100 mg/l endotoxin for 96 h no adverse effects were observed. In a similar test, embryos and larvae were more sensitive with LC50 values ranging from 86 to 188 mg/l for the four toxins; behavioral disturbances as well as delayed or abnormal development were also observed, particularly for Cry1Aa. When the adults were examined for the presence of abnormal micronuclei formation in peripheral blood erythrocytes, Cry1Aa alone or in combination either with Cry1Ac or

Cry2A significantly increased the frequency of micronuclei. Interestingly, antagonism was observed for the Cry1Aa  Cry1Ab combination. This study demonstrates that toxicological evaluation of Bt toxins should be done according to actual field exposures and should consider possible interactions with other natural or synthetic toxicants. Tayabali and Seligy (2000) studied in great detail the cytolytic effect of B. cereus, Bt subsp. israelensis and Bt subsp. kurstaki proteinaceous culture components on several mammalian cells, including human colonic epithelial, liver, and blood cell lines in vitro. The main findings of the study are as follows. During their vegetative growth, all three bacteria produced a range of thermolabile exotoxins (VCPs), which were present in the supernatant fractions of the culture. Above threshold concentrations, the VCPs showed nonspecific toxicity to mammalian cells, as indicated by loss in bioreduction, cell rounding, blebbing, protein degradation and cytolysis. Furthermore, overall protein synthesis in the exposed mammalian cells was inhibited, presumably by ADP-ribosylating toxins in the bacterial culture. The presence of the antibiotic gentamicin in the culture suppressed the production of the cytotoxic products. Experiments also confirmed that -endotoxins or related/derived proteins are not toxic to human cells. Poisoning symptoms, biochemistry, and pathology  The solubilized Bt subsp. israelensis endotoxin crystals caused poisoning symptoms and death in mice receiving 16.6 or

Chapter | 3  Pest Control Agents from Natural Products

33.3 mg/kg doses on intravenous but not on oral administration (Thomas and Ellar, 1983). Within 1 h of injection, the animals developed paralysis in their hindquarters and became relatively immobile within 3 h. Breathing and heart rate increased. Death occurred after 12 h or 36–48 h after injection at the higher and lower doses, respectively. In mice, solubilized endotoxin and immunoaffinitypurified toxin fractions, a 28-kDa protein fraction in particular, of Bt subsp. israelensis given at 10, 7.5 and 2.5 mg/kg respective intraperitoneal doses caused hypothermia and bradycardia (Mayes et al., 1989). Cytolysis of red and white blood cells was not detectable after a 2.5-mg/kg intraperitoneal dose of solubilized endotoxin. Pathological and histological examinations of rats and mice treated intraperitoneally with 5 mg/kg solubilized endotoxin revealed focal-to-segmental reddened and edematous areas within the small intestine with major lesions in the jejunum. The toxic symptoms in the Japanese quail receiving larger than 10 mg/kg intraabdominal doses of soluble Bt subsp. israelensis endotoxin were loss of alertness, loss of activity in the legs, and loss of substantial volumes of fluid in the cloacal excreta within 2 h of injection (Kallapur et al., 1992). Bradycardia and hypothermia were observed for the 30-mg/kg treatment. This dose reduced serum lipid and alkaline phosphate levels while it increased serum glucose, creatine phosphokinase, and lactate dehydrogenase. Serum calcium, alanine transaminase, blood urea nitrogen, bilirubin, and protein levels were the same as for the control. Toxicity of -exotoxin  -Exotoxin showed delayed toxicity to mammals. The pure substance obtained from Bt subsp. gelechiae gave an LD50 of 18 mg/kg in mice, as estimated on the third day after intraperitoneal application (Šebesta et al., 1969). The exotoxin is much less toxic to mice when given orally and upon dephosphorylation toxicity is lost completely (see Šebesta et al., 1981). -Exotoxin, obtained from the culture supernatant of Bt subsp. morrisoni, gave subcutaneous LD50 values of 184.5 and 135.6 mg/kg for male and female mice, respectively (Haufler and Kunz, 1985). The oral LD50 of purified -exotoxin was about 170 mg/kg in rats and caused dermal toxicity to rabbits at 0.4 mg/kg (see McClintock et al., 1995). High doses of Bt formulations containing -exotoxin are toxic to honeybees (Vandenberg and Shimanuki, 1986). In plant cells, the toxin inhibited mitosis and impaired microtubules of the spindle and the phragmoplast, effects characteristic to colchicine and vinblastine (Sharma and Sahu, 1977). (h)  Toxicity to Humans Allergic reactions, infection and toxicity problems with Bt can arise during manufacture (fermentation), handling, and field use. Physical and laboratory examination of human volunteers who inhaled 3  108 viable spores of Bt Berliner as powder for 5 days and ingested 3  109 viable

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spores of the bacterium daily for 5 days revealed no genitourinary, gastrointestinal, cardiorespiratory, or nervous system anomalies (Fisher and Rosner, 1959). Oral and dermal administrations of 105–109 cells/g of Bt var. galleriae preparations produced nausea, belching, vomiting, tenesmus, colic-like pain, diarrhea, and fever, symptoms similar to those caused by B. cereus food poisoning (Pivovarov et al., 1977) but these effects could have been related to the -exotoxin present in the preparation (see Ray, 1991). Volunteers receiving 1  1010 Bt spores daily for 3 days showed no treatment-related symptoms, and viable Bt spores could be recovered from half of the patients for 30 days after treatment. Frederiksen et al. (2006) analyzed fresh fruit and vegetable products sold in Danish retail shops for endotoxin crystals, cry genes, and enterotoxigenic profile. Among the 128 isolates, mostly from pepper, cucumber, and tomatoes, 14 contained bacterial strains indistinguishable from Bt subsp. kurstaki while nine strains contained Bt subsp. aizawai, components of Bt-based insecticides used in greenhouses in Denmark and The Netherlands, respectively. The authors warn about the risk of gastroenteritis due to the presence of Bt strains harboring genes for hemolysin BL toxin, nonhemolytic enterotoxin, and cytotoxin K. Similarly, Zhou et al. (2008) have recently reported Bt subsp. kurstaki insecticide residues in milk products and tea beverages. Irritation and sensitization  Two reported incidents of possible allergic reactions to Bt-based products were unrelated to the bacterium but likely caused either by a previously undiagnosed disease (Kawasaki syndrome) or by an existing food allergy elicited by a formulating ingredient (see McClintock et al., 1995; Siegel, 2001). A recent study examining the allergenic potential of farm workers exposed to various levels of Bt insecticide revealed no adverse respiratory symptoms, such as asthma, but skin allergy responses lasting up to 4 months and elevated levels of IgG and IgE antibodies were found especially in high-exposure groups (Bernstein et al., 1999). Examination of the fecal samples of 20 greenhouse workers exposed to Bt subsp. israelensis revealed that the isolates of seven workers contained parasporal crystal toxins, the genes for cry11 as well as for hemolytic and nonhemolytic enterotoxins (Jensen et al., 2002). In these persons, enterotoxin production was also evident and the majority of the excreted cells were in the vegetative state suggesting that the ingested spores could have germinated in the gastrointestinal tract. However, manifested gastrointestinal symptoms did not correlate with the presence of the toxin. Poisoning incidents  The first reported occurrence of an infection caused by Bt in humans was due to the accidental splashing of Dipel insecticide suspension in the eye of a

162

farmer (Samples and Buettner, 1983a, 1983b). In spite of immediate rinsing with water and application of antibiotic ointment, the eye was still irritated 3 days later when local corticosteroid treatment was begun. Ten days after the accident, a corneal ulcer was noted. The bacterium could be cultured from the eye and proved to be susceptible to gentamicin, which cured the patient. Another case of infection occurred in a laboratory when a student working with spores and endotoxin of Bt subsp. israelensis and Acinetobacter calcoaceticus var. anitratus accidentally stuck his finger on a needle (Warren et al., 1984). Within 2 h, the finger became painful then discolored and the hand was swollen. In spite of immediate antibiotic therapy consisting of intravenous gentamicin first then benzylpenicillin (2.5 g every 4 h), lymphangitis developed. After 24 h, the flexor tendon sheath required decompression over the finger joint close to the inoculation site. The patient recovered after 5 days. It was also discovered that the crystalline israelensis -endotoxin protoxin could be activated in vitro at room temperature or 30°C within 2–4 h by proteases present either in Bt or in culture filtrates from unrelated bacteria, including the A. calcoaceticus var. anitratus (see also Damgaard et al., 1997). Green et al. (1990) described the results of an epidemiological study conducted in connection with an extensive Bt spraying program to control gypsy moth in Oregon. Of the 55 cultures from human specimens positive for Bt, only four cases could be related to the insecticide treatment. One of these was a sprayer who accidentally splashed the Bt spray mixture on his face and eyes and developed dermatitis, pruritis, burning, swelling, and erythema, with conjunctival injection. The eyelid and skin were treated with steroid cream. Bt was cultured from 18 different body sites or fluids from the other 54 cases, suggesting that the bacterium was appearing as a contaminant or commensal, rather than a pathogen, because there was no consistent pattern of disease associated with its presence. The authors also pointed out that immunocompromised persons could be at risk when exposed to Bt-based insecticides. Otvos et al. (2007) reviewed related results of a series of epidemiological studies carried out across Canada during spruce budworm and gypsy moth eradication campaigns using Bt subsp. kurstaki between 1984 and 2000. Occupational exposure typically ranging from 106 to 108 CFU/m3 caused transient spore contamination in 3–30% of the personnel as shown by immunological or bacterium culture analyses. Workers exposed to higher concentrations retained Bt at least for 5–6 days, and most were culture positive for 14–30 days. The most common symptoms were eye, nose and throat irritation, dry skin and chapped lips; complaints were most prevalent among workers with histories of allergies. Nonoccupational exposure to the low spore concentrations (less than 103 CFU/m3 on the day of spray) of airborne Bt represented minimal hazard for the population in the treated areas.

Hayes’ Handbook of Pesticide Toxicology

Treatment  Treatment is symptomatic. Inflammations (dermatitis, erythema) can be treated with steroids while infections require antibiotics (see, for example, Luna et al., 2007). Bt is known to be resistant to some (penicillin and ampicillin) but not all -lactam antibiotics. An opportunistic Bt superinfection of a wounded soldier could be treated with a 10-day regimen of ciprofloxacin and gentamicin (Hernandez et al., 1998a).

3.2.2.2  Spinosad (a)  Introduction Spinosad is a novel selective insecticide containing structurally unique glycosylated macrolactones (spinosyns) with activity against a broad range of insect pests, including important lepidopteran and dipteran species. Its insecticidal activity was discovered in the mid-1980s (reviewed by DeAmicis et al., 1997; Kirst et al., 1992; Salgado and Sparks, 2005; Thompson et al., 2000). Spinosyns, initially referred to as A83543 factors, are isolated from the fermentation broth of the aerobic, Gram-positive soil bacterium Saccharopolyspora spinosa (Actinomycetes). More than 20 different spinosyns have been identified from wild-type and mutant S. spinosa cultures. Cloning and sequencing the spinosyn biosynthetic gene cluster have been used to clarify the pathways involved in these unique macrolides (Waldron et al., 2001; for a review on spinosyn biosynthesis, see Huang et al., 2009). Furthermore, extensive structure– activity studies with the natural products and their derivatives (spinosoids) have culminated in the development of spinetoram (XDE-175), a semisynthetic derivative with improved insecticidal properties (Crouse et al., 2007). Recently, homologous compounds, the butenylspinosyns (pogonins), have been isolated from the culture of a related actinomycete, Saccharopolyspora pogona (Hahn et al., 2006). (b)  Identity, Physicochemical Properties, and Uses Spinosad consists of about 70–85% spinosyn A and 15–30% spinosyn D (Figure 3.8) with traces of structurally related components. Spinosyn A IUPAC name: (2R,3aS,5aR,5bS,9S,13S,14R,16aS,1 6bR)-2-(6-deoxy-2,3,4-tri-O-methyl--l-mannopyranosyloxy)-13-(4-dimethylamino-2,3,4,6-tetradeoxy--d-erythropyranosyloxy)-9-ethyl-2,3, 3a,5a,5b,6,7,9,10,11,12,13,14,15,16a,16b-hexadecahydro-14-methyl-1H-as-indaceno[3,2-d]oxacyclododecine7,15-dione. Chemical Abstract name: (2R,3aS,5aR,5bS,9S,13S,14R, 16aS,16bR)-2-[(6-deoxy-2,3,4-tri-O-methyl--Lmannopyranosyl)oxy]-13-[[(2R,5S,6R)-5-(dimethylamino)tetrahydro-6-methyl-2H-pyran-2-yl]oxy]-9-ethyl-2,

Chapter | 3  Pest Control Agents from Natural Products

N

dichloromethane, methanol and toluene; spinosyn D is somewhat less soluble in these solvents.

O

O

O O H

O O O

H

O

O

O

H

163

H H R

spinosyn A R = H spinosyn D R = CH3

Figure 3.8  Structures of the components of spinosad insecticide.

3,3a,5a,5b,9,10,11,12,13,14,16a,6b-tetradecahydro-14methyl-1H-as-indaceno[3,2-d]oxacyclododecin-7,15dione. CAS Registry Number: [131929-60-7]. Empirical formula: C41H65NO10; molecular weight: 732.0. Spinosyn D  IUPAC name: (2S,3aR,5aS,5bS,9S,13S,14R, 16aS,16bS)-2-(6-deoxy-2,3,4-tri-O-methyl--lmannopyranosyloxy)-13-(4-dimethylamino-2, 3,4,6-tetradeoxy--d-erythropyranosyloxy)-9-ethyl-2,3, 3a,5a,5b,6,7,9,10,11,12,13,14,15,16a,16b-hexadecahydro-4,14-dimethyl-1H-as-indaceno[3,2-d]oxacyclododecine-7,15-dione. Chemical Abstract name: (2S,3aR,5aS,5bS,9S,13S,14R, 16aS,16bS)-2-[(6-deoxy-2,3,4-tri-O-methyl--lmannopyranosyl)oxy]-13-[[(2R,5S,6R)-5-(dimethylamino) tetrahydro-6-methyl-2H-pyran-2-yl]oxy]-9ethyl-2,3,3a,5a,5b,6,9,10,11,12,13,14,16a, 16b-tetradecahydro-4,14-dimethyl-1-H-as-indaceno[3,2-d]oxacyclododecin-7,15-dione. CAS Registry Number: [131929-63-0]. Empirical formula: C42H67NO10; molecular weight: 746.0. Physicochemical properties  The physicochemical properties of the two main active ingredients in spinosad are somewhat different (DeAmicis et al., 1997). The white-gray crystals of pure spinosyn A melt at 118–124°C. The solubility in water is 290, 235, and 16 mg/ kg at pH 5.0, 7.0, and 9.0, respectively; at pH 5.0 and 7.0 the log P values are 2.8 and 4.0, respectively. The pKa of spinosyn A is 8.1. Optical rotation: []D  135.3° (c  1.0 in ethanol). The white–gray crystals of pure spinosyn D forms melt at 169–174°C. The solubility in water is 29, 0.332, and 0.053 mg/kg at pH 5.0, 7.0, and 9.0, respectively; at pH 5.0 and 7.0 the log P values are 3.2 and 4.5, respectively. The pKa of spinosyn D is 7.87. Optical rotation: []D  156.7° (c  1.0 in ethanol). Spinosyn A is slightly to moderately soluble in polar organic solvents, such as acetone, acetonitrile, ethyl acetate,

Formulations and uses  Commercial spinosad is a mixture of two active components, spinosyn A and spinosyn D. It is mainly sold as a water-based suspension concentrate or as water-dispersible granules. It is currently used to control lepidopteran, as well as some coleopteran, dipteran and thysanopteran pests on cotton, vegetables, orchards, coffee, turfgrass and ornamentals at rates of 5–150 g/ha. Spinosad is also efficacious against thrips in citrus and against stored products pests and also used in (fire) ant control. Novel, highly selective “attract-and-kill” baits containing spinosad as the insecticidal component have been developed for fruit fly control (see, for example, Mangan et al., 2006; Vargas et al., 2008) and mosquitoes (Müller et al., 2008). Kirst et al. (2002) reviewed spinosad products that control parasitic pests in mammals. Recently, De Deken et al. (2004) have demonstrated the effectiveness of spinosad in tsetse fly control. Beef-flavored, chewable spinosad tablets have been commercialized to control fleas on dogs (Robertson-Plouch et al., 2008). Stability and residues  Both commercial spinosyns are stable in solutions with pH values between 5 and 9, thus abiotic hydrolysis in the environment is relatively unimportant. Decomposition occurs in strong acidic or basic solutions. The aqueous solution of the technical-grade material is stable in the dark and has a pH value of 7.74. Under natural light conditions, however, the photolytic half-life of spinosyns A and D in aqueous solution (pH  7.0 and 25°C) is less than 1 day; for the formulated product in a natural water microcosm half-lives of up to 2 days were observed (Cleveland et al., 2002). Liu and Li (2004) noted that the photodegradation of spinosad was faster in tap water, seawater and stream water than in distilled water, indicating that solutes as photosensitizers may accelerate decomposition of spinosyns. The half-life of spinosad in soil is about 2 weeks (Saunders and Bret, 1997). In a forest environment Thompson et al. (2002a) observed faster decomposition: the half-life of spinosyn A in litter and open soil was 2.0 and 7.8 days, respectively; spinosyn D dissipated more rapidly. Under subtropical conditions the half-lives of spinosad in soil and vegetables were 1–3 days with residues detectable a week after treatment (Sharma et al., 2007; Zhao et al., 2007). Spinosad, is much more stable indoors; for example, it retained full insecticidal activity for 9–12 months in stored wheat (Daglish and Nayak, 2006; Fang et al., 2002) or tobacco (Blanc et al., 2004). Furthermore, spinosad in spray and pour-on formulations provided complete control of chewing and sucking lice on cattle for 7–8 weeks (White et al., 2007).

164

Rutherford et al. (2000) determined spinosad residues in milk and various tissues after chronic oral dosages to dairy cows. After feeding the animals with diets containing up to 10 ppm spinosad for 28 days, the relatively lipophilic insecticide accumulated in fat (5.7 g/g), milk cream (1.9 g/g), and liver (1.2 g/g), while skim milk, kidney, and lean beef contained less than1 g/g at the maximal dosage rate. In a related experiment with laying hens fed up to 5 ppm dietary spinosad for 42 days, the maximum dosage rate resulted in 0.19 and 0.17 g/g spinosad residues in the eggs and whole body, respectively, while the residue in abdominal fat was as high as 1.2 g/g. A more recent study by Rothwell et al. (2005) with sheep sprayed spinosad spray showed that the highest residue level (0.2 g/g) was in fat. A monograph describing the results of a number of studies on the environmental fate and residues of spinosad under various conditions and applications is available (World Health Organization and Food and Agriculture Organization, 2002). (c)  Biological Properties Mode of action  Spinosad kills insects by contact or ingestion with a speed comparable to most neurotoxic insecticides. It is not systemic but can slowly penetrate leaves. Based on poisoning symptoms, electrophysiological and receptor studies using spinosyn A, the main mode of insecticidal action of these macrocyclic lactones has been clarified (reviewed by Salgado and Sparks, 2005; see also Millar and Denholm, 2007). The poisoning symptoms caused by spinosyns are due to their CNS-stimulatory action and can be divided into three phases (Salgado, 1998; Salgado et al., 1998). For the adult American cockroach, Periplaneta americana, following the injection of an LD50 dose (1.9 g per animal) the first symptoms are prolonged involuntary muscle contractions leading to postural changes, typically elevation of the body and straightening of the hindlegs. In the second phase of poisoning, after many hours of hyperexcitation, incoordinated movement ensues, fine tremors appear in all muscles, and the insect falls on its back. In the final phase, apparently due to neuromuscular fatigue, all movements and tremors cease and paralysis follows. In electrophysiological studies with cockroach neurons, 20 nM spinosyn A caused depolarization and increased the spontaneous firing rate. These effects could be reversed or blocked by the selective nAChR antagonists methyllycaconitine or -bungarotoxin. In receptor studies, spinosyn A and competitive nAChR agonists such as imidacloprid were shown to bind to nonoverlapping, possibly allosteric binding sites or/and at separate receptor subtypes (Salgado and Saar, 2004). In addition to acting as agonists of ligand-gated nAChRs, spinosyns antagonized GABA-activated chloride channels in isolated cockroach neurons at nanomolar concentrations

Hayes’ Handbook of Pesticide Toxicology

(Watson, 2001) and this effect is likely to contribute to the overall toxicity of the insecticide. In contrast to conventional aminosugar macrolide derivative drugs, such as erythromycin, spinosyns are devoid of antibacterial activity. Metabolism and excretion  Studies on the metabolism and fate of spinosyn A and D have been reviewed (Salgado and Sparks, 2005; see also World Health Organization, 2002; World Health Organization and Food and Agriculture Organization, 2002). Insects, in general, appear to have limited capacity to metabolize spinosyn A. In the cockroach Periplaneta americana, over 60% of radiolabeled spinosyn A was metabolized by 64 h into more polar products, which have not been characterized, but presumably result from N- and O-demethylation. While the resistance to spinosad has most often been associated with altered nAChR, other mechanism(s) might also be operating (reviewed by Scott, 2008). Interestingly, resistance observed in a field population of the diamondback moth could be partially overcome by PB and S,S,Stributyl phosphorotrithionate synergists indicating the involvement of enhanced oxidative demethylation and ester (lactone) hydrolysis in this population; these synergists did not increase the toxicity of the insecticide in nonresistant strains of this moth (Sayyed et al., 2008). Metabolism studies with 14C-labeled spinosyns in rats showed that the absorption, distribution, metabolism and elimination of spinosyns A and D were similar. For spinosyn A, 28 and 52% of the administered radioactivity was eliminated in the feces of females and males, respectively, while a small amount of the applied radioactivity was eliminated in the urine. Among the metabolites, N- and O-demethylated products were identified. Tentative cysteine or glutathione conjugates were also isolated. The elimination half-lives of spinosyns A and D were 25–42 h and 29–33 h, respectively. In lactating goats, about two-thirds of the radiolabeled spinosyns, dosed at 10 ppm in the diet for 3 days, was excreted with the feces by the fourth day. Of the radioactive residue found in milk, spinosyns A and D accounted for 1.8 and 0.69% of their respective ingested total radioactivity. In fatty tissues, 33 and 17% of the respective total dose of spinosyns A and D was found. In addition to the N- and O-demethylated metabolites observed in rats, compounds hydroxylated on the macrolide ring could be partially characterized. In addition to residue studies discussed above, uptake, distribution and metabolism of spinosad has been examined in some fruits, vegetables and cotton using 14C-labeled spinosyns A and D. In general, most of the insecticide sprayed onto plants remains on the surface for several days although in gradually decreasing amounts due to photodegradation and, to some extent, rain (World Health

Chapter | 3  Pest Control Agents from Natural Products

Organization and Food and Agriculture Organization, 2002). In experiments with apple trees, the total radioactivity on the fruit decreased by about 50% over 6 weeks. Some residue penetrated into the peel and the pulp but over 60% could be rinsed from the surface. Among the polar metabolites, N-demethylated products could be identified. The fate of spinosad on grape was similar. On tomato the dissipation of [14C]spinosyn A from the surface of the plant occurred faster (only 24% could be recovered 3 days after treatment). In turnips, most of the applied spinosyns transformed into demethylated metabolites and other (photo)degradation products in the foliage within 10 days after treatment; however, a transient increase in the root was seen: for example, the amount of spinosyn A in the foliage and root was 0.066 and 0.084 mg/kg, respectively, 24 days after treatment and this relative difference was more pronounced 48 days after treatment when the foliage and root contained 0.003 and 0.047 mg/kg spinosyn A, respectively. In cotton, analyses 6 weeks after treating the plants with radiolabeled spinosad indicated extensive metabolism (degradation) of the parent compounds into unidentifiable metabolites (degradates) and the incorporation of the radiolabel into cellulose and/or fatty acid oil components could be demonstrated. Finally, it must be noted that information on the biological profile of the individual metabolites or of the persistent environmental degradates is currently scarce. (d)  Toxicity to Animals and Humans Acute and chronic toxicity to animals  The toxic effects of spinosad on a broad range of experimental animals and nontarget organisms have been studied. In general, spinosad has low acute mammalian toxicity. Relevant toxicity data for spinosyn A are summarized in Table 3.8 (see also Cleveland et al., 2001; World Health Organization, 2002). Due to the low risks it presents to humans and the environment, spinosad is classified by the U.S. EPA as a reduced-risk material. In acute and subchronic tests, spinosad did not demonstrate any neurotoxic or reproductive effects in rats, mice, and dogs (reviewed by World Health Organization, 2002). Recent publications have disclosed the details of subchronic and chronic toxicity studies with dietary dosages of spinosad for mice (Stebbins et al., 2002) and rats (Yano et al., 2002). In a 13-week study, mice were provided with diets containing 0–0.12% spinosad, consisting of either 77% spinosyn A and 23% spinosyn D, or 88% spinosyn A and 12% spinosyn D. In an 18-month chronic study, the diets contained 0 to 0.036% spinosad. At doses 0.015% of spinosad the primary toxic effects were intracellular vacuolation of histiocytic and epithelial cells in numerous tissues and organs. Cardiac and skeletal muscle cells were also vacuolated, although to a lesser degree. These effects, unprecedented among pesticides, are typically associated with phospholipidosis, a condition that results from

165

accumulation of polar lipids in lysosomes and is induced by various cationic amphiphilic drugs (Reasor et al., 2006). Further histological examination of animals that received high dosages (generally above 0.024%) noted degenerative changes in the kidneys, necrosis of the liver and bone marrow, hyperplasia of the glandular mucosa of the stomach, and anemia associated with hematopoiesis in the spleen. Furthermore, at 0.12% spinosad concentration, corresponding to 109.7 mg/kg for males and 141.9 mg/kg for females daily doses, overt toxic effects, weight loss likely due to severe gastric hyperplasia, and 25% mortality mainly due to hepatic necrosis were observed. For either sex, the dietary NOEL in the 13-week study was 0.005% (6 mg/ kg daily dose) spinosad, in the chronic study the dietary NOEL was 0.008% (11 mg/kg daily dose) spinosad. In a 13-week subchronic toxicity study with rats given feed containing 0–0.4% spinosad preparations, consisting of spinosyns A and D at either 5.5:1 or 6.4:1 ratio, respectively, Yano et al. (2002) observed similar effects as for the mouse study described above. Specifically, at 0.05% spinosad concentrations cytoplasmic vacuolation was present in various tissues; at 0.2% spinosad concentrations, degenerative or regenerative changes in tissues of the liver, skeletal muscles, testes and the stomach. Inflammation was also observed in some tissues. In the 2-year oncogenicity study, vacuolation and inflammation was also seen in the thyroid glands in rats given 0.05% spinosad for 1 year; by 21 months, excessive mortality occurred in rats for spinosad at 0.1% dietary concentrations. Significant toxic effects (thin appearance, deep and labored respiration, and hyperthermia) as well as mortality were observed at 0.4% spinosad concentration. The NOEL in the 13-week study was 0.012% dietary spinosad (24 mg/kg daily dose). In the 2-year study, the NOEL was 0.005% dietary spinosad (2.4 mg/kg daily dose). Comparing these data with those obtained in earlier (unpublished) experiments that used a 1:1 mixture of spinosyns A and D, the authors concluded that changes in the ratio of the two spinosyns did not significantly affect the mammalian toxicity of various spinosad compositions. Reviewing the available literature on the laboratory and field effects of spinosad on 52 species of insect natural enemies, Williams et al. (2003) found that predators do not suffer significant sublethal effects following exposure to spinosad, although there were exceptions, for example earwigs (see also Van Driesche et al., 2006). By contrast, parasitoids are generally more susceptible but with judicious agricultural techniques exposure can be minimized (Ruiz et al., 2008). According to U.S. EPA classification for pollinators, spinosad is “highly toxic,” based on an acute 48-h topical LD50 of 2.5 ng per bee (Bret et al., 1997), yet the use of spinosad carries significantly less ecotoxicological risk than most other insecticides, especially when considering its photolability (Mayes et al., 2003).

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166

Table 3.8  Acute Toxicity of Spinosad Species

Route

LD50 (mg/kg)a

Rat, male

oral

3738

Bret et al. (1997)

Rat, female

oral

5000

Bret et al. (1997)

Mouse

oral

5000

Bret et al. (1997)

Rabbit

dermal

5000

Bret et al. (1997)

Rat

inhalation

LC50  5.18 mg/l air

Bret et al. (1997)

Rabbit

eye

Slight conjunctival irritation, clearing in 48 h

Bret et al. (1997)

Guinea pig

dermal

No sensitization

Bret et al. (1997)

Rabbit

dermal

No irritation

Bret et al. (1997)

Bobwhite quail

oral

Bobwhite quail

5-day dietary

Mallard duck

oral

Mallard duck

5-day dietary

LC50  5000 ppm

Bret et al. (1997)

Rainbow trout

96-h

LC50  30 ppm

Bret et al. (1997)

Bluegill

96-h

LC50  5.9 ppm

Bret et al. (1997)

Carp

96-h

LC50  5.0 ppm

Bret et al. (1997)

Sheepshead minnow

96-h

LC50  7.9 ppm

Bret et al. (1997)

Coho salmon

Other test results

References

Bret et al. (1997)

2000

Bret et al. (1997)

LC50  5000 ppm

Bret et al. (1997)

2000

96-h

LC50  500 ppm

Deardoff and Stark (2009)

b

48-h

LC50  92.7 ppm

Bret et al. (1997)

b

Daphina magna

48-h, static

EC50  1.0 ppm

World Health Organization (2008)

Daphina magnab

21-day, flowthrough

NOEC  1.2 ppbc

World Health Organization (2008)

Daphina magnab

48-h, static

NOEC  6.7 ppbc

Daphina magna

b

Daphina magna

b

d

21-day, static

NOEC  6.88 ppb

Duchet et al. (2008) c

Cleveland et al. (2001) c

Daphina magna

21-day, flowthrough

NOEC  0.617 ppb

Daphina magnab

21-day, semi-static

NOEC  8.0 ppbc

Cleveland et al. (2001)

Daphina magnab

48-h

LC50  4.8 ppb

Deardoff and Stark (2009)

Daphina pulex

48-h

LC50  129 ppb

Deardoff and Stark (2009)

Green alga, Selenastrum capricornutum

7-day

EC50  105 ppm

Bret et al. (1997)

Blue green alga, Anabaena flosaquae

5-day

EC50  8.1 ppm

Bret et al. (1997)

Grass shrimp

96-h

LC50  9.8 ppm

Bret et al. (1997)

Honeybee

topical, 48-h

(LD50  2.5  ng/bee)

Bret et al. (1997)

Honeybee

oral

(LD50  60  ng/bee)

Cleveland et al. (2001)

a

Cleveland et al. (2001)

Values are in mg/kg unless otherwise noted (in parenthesis). b The reported large variation in the toxicity of the photodegradable spinosad to Daphnia species is due to different experimental conditions (static or flow-through system; different absorbed oxygen content, etc.), to the age-dependent sensitivities during the recovery of the water flea population, and also to some unidentified factors, possibly the toxicity of the degradation products. (For a discussion, see Cleveland et al., 2001; Duchet et al., 2008; National Registration Authority for Agricultural and Veterinary Chemicals, 1998.) c No (adverse) effect concentration. d During the 21-day test, 5-day pulses of spinosad exposures were used that were followed by dilutions to simulate natural exposure scenarios.

Chapter | 3  Pest Control Agents from Natural Products

Effect on reproduction, carcinogenicity, and mutagenicity  Breslin et al. (2000) evaluated the maternal and embryonal/fetal toxicity of spinosad in rats and rabbits. The insecticide was administered by gavage daily to pregnant rats at 0–200 mg/kg and to pregnant rabbits at 0–50 mg/kg on gestation days 6–15 and 7–19, respectively. At the end of the experiments (at the 21st gestation day for rats and the 28th gestation day for rabbits) the animals were evaluated for maternal organ weight, reproductive parameters, fetal body weight, and fetal external, visceral and skeletal muscle structures. In general, spinosad treatment did not cause any significant differences in terminal maternal body weights or relative organ weights, but due to decreased food consumption dose-related transient body weight losses were observed both in rats and rabbits during the initial part of the treatment period. Furthermore, no treatment-related embryonal or fetal toxicity was seen in the animals at any spinosad dose tested. The NOELs for maternal toxicity were daily spinosad doses of 50 and 10  mg/kg for rats and rabbits, respectively. The embryonal/fetal NOELs were 200 and 50 mg/kg daily doses for rats and rabbits, respectively. The effect of spinosad on reproduction was examined in a two-generation dietary study using rats given diets that provided 0–0.2% spinosad concentration in the diet corresponding to approximate daily spinosad doses of 0–100 mg/kg body weight (Hanley et al., 2002). The effect of the insecticide was evaluated by recording body weight changes and reproductive performance, and by pathological examinations. Following the 10-week (P1 generation) or 12-week (P2 generation) dietary exposures, the 100 mg/kg daily dose depressed body weight gains by 7 and 3.5% relative to control for the P1 and P2 male generations, respectively; females were generally less affected. No such difference was seen at either sex at the lower dosages. The 100-mg/kg daily spinosad dose produced parental toxicity and affected the offspring. Absolute and relative liver, kidney, heart, spleen, and thyroid weights were increased by 12–240% of control values. The histological changes were consistent with findings of the subchronic and chronic studies described above (Yano et al., 2002) and included vacuolation (especially in the thyroid), aggregates of macrophages/reticuloendothelial cells, degenerative and/or inflammatory responses. Notably, the highest spinosad dose produced chronic inflammation of the prostate in males of both generations but no other histological changes were found in reproductive organs of either males or females. No treatment-related histological changes were observed at the lower doses. Spinosad treatment did not affect male or female mating indices and no statistically significant differences could be seen for the fertility indices and gestation length. Treatment-related effects on the offspring were observed at the 100 mg/kg daily dose only; for example, the percentage of pups born alive was lower in the treated animals suggesting early neonatal effects secondary to the effects on maternal animals. The NOEL

167

for both parental and reproductive/perinatal toxicity was 30 mg/kg daily dose. Based on the results, spinosad was not considered a selective reproductive toxicant. In chronic oncogenicity studies with mice (Stebbins et al., 2002) or rats (Yano et al., 2002), no increase in the incidence of tumors relative to controls were observed at dietary dose levels up to 0.036% spinosad for mice or 0.05% spinosad for rats. A mutagenic battery consisting of the Chinese hamster ovary, mouse lymphoma cell, mouse bone marrow micronucleus, rat hepatocyte unscheduled DNA synthesis, and Ames tests did not show mutagenic activity for the insecticide (World Health Organization, 2002, 2008). (e)  Toxic Effects in Humans There have been no described cases of severe human poisoning attributable solely to spinosad use (see, for example, Calvert et al., 2008). Dietary risk assessments of spinosad have been published (Gao et al., 2007; World Health Organization and Food and Agriculture Organization, 2002).

3.2.3  Semiochemicals In principle, semiochemicals are natural products involved in animal communication. These behavior-modifying compounds can be divided into two main groups, pheromones and allelochemicals. Pheromones mediate communication between individuals of the same species, whereas allelochemicals act between different species. The latter can be further divided into allomones, which give advantage to the emitter (e.g. defensive secretions), and kairomones, which give advantage to the receiver (e.g. secretions that can be detected by predators or parasites). This section will deal only with two types of semiochemicals: (1) insect sex pheromones, which are volatile compounds indispensable in mate finding, and (2) kairomones, which are important cues in locating hosts on which the insect feeds.

3.2.3.1  Insect Sex Pheromones (a) Introduction and General Properties Insect sex pheromones are emitted by one sex, usually females, of a particular species in miniscule amounts, and are perceived by members of the opposite sex, eliciting complex behavioral responses including mate searching and mating. Sex pheromones are highly species specific and, as a rule, a unique blend of the natural pheromone components is needed for attraction. The first sex pheromone identified was (10E,12Z)-10,12-hexadecadienol or bombykol, the sex pheromone of the silkworm moth, Bombyx mori (Butenandt et al., 1959). Since then, pheromones for over 500 insect species have been identified and introduced into agriculture in one form or another to

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complement or, in some cases, replace conventional pest control agents. There is also a large number of attractants discovered by empirical screening of natural and synthetic chemicals. Structurally, the sex pheromones encompass a diverse group of volatile, predominantly lipid-like straight chain aliphatic alcohols, esters or aldehydes, but several terpenoid or polycyclic oxygenated structures have also been identified. Structures of selected sex and aggregation pheromones are depicted in Figure 3.9. Over 2000 pheromonal compounds have been isolated and identified from volatile secretions of insect pests and disease vectors. The subject, including the chemistry, biosynthesis, physiological and behavioral aspects, mode of action and application, was thoroughly reviewed (Ando et al., 2004; Baker and Heath, 2005; Blomquist and Vogt, 2003; Rafaeli, 2005). Pheromones reported until 1988 were compiled by Mayer and McLaughlin (1991); up-to-date

lists of known pheromones, attractants, and related semiochemicals are available on the Internet (El-Sayed, 2008; Witzgall et al., 2004). (b)  Uses and Formulations Pheromones are applied in four major ways (Cork, 2004; El-Sayed et al., 2006; Howse et al., 1998; Jutsum and Gordon, 1989; Villavaso et al., 2003): (1) population monitoring with traps baited with the pheromone; (2) mass trapping using a large number of high-capacity trapping devices; (3) pheromone plus insecticide combination (lureand-kill); and (4) mating disruption by permeating the area with specially formulated pheromones that are then released over several weeks into the air. For economic reasons, it is not the naturally produced pheromone but its synthetic equivalent that is used in practice. The amount of the pheromone blend in traps generally ranges from 0.1 to 1000 mg, whereas mating disruption O O

muscalure gossyplure H

n

O

CHO

H (+)-disparlure

(Z)-11-hexadecenal n = 2 (Z)-13-octadecenal n = 3

R

OH

H grandlure I (Z), R = CHO grandlure II (E), R = CHO grandlure III (Z), R = CH2OH

grandlure IV (grandisol)

OH codlemone

O

HO

HO

O

HO H

ipsdienol

ipsenol

cis-verbenol

frontalin

Figure 3.9  Structures of selected insect pheromones. Muscalure, the sex pheromone of the housefly, Musca domestica; gossyplure, the sex pheromone of Pectinophora gossypiella; ()-disparlure, the natural stereoisomer of the sex pheromone of the gypsy moth, Lymantria dispar; (Z)-11-hexadecenal and (Z)-13-octadecenal, common sex pheromone components of Lepidoptera, including the rice stem borer, Chilo suppressalis; grandlure I, II, III, and IV, the sex pheromone components of the boll weevil, Anthonomus grandis; codlemone, the sex pheromone of the codling moth, Cydia pomonella; ipsdienol, ipsenol, and cis-verbenol, the aggregation pheromone blend of the bark beetle, Ips paraconfusus; and frontalin, an aggregation pheromone component of Dendroctonus bark beetles.

Chapter | 3  Pest Control Agents from Natural Products

typically requires 50–200 g/ha of the pheromone to provide the necessary aerial concentration, typically 1–20 ng/m3 (see, for example, Flint et al., 1993; Koch et al., 2002), for extended periods. Species-selective traps for population monitoring and detection are available for hundreds of pest insects worldwide. Mass trapping and mating disruption have also been used in cotton, orchards, grape, vegetables, and forestry. (c)  Stability Because of their particular mode of action, pheromones are volatile substances and require special formulations providing even emission of the pheromone blend for weeks. Pheromones containing double or triple bonds, especially conjugated ones, and/or aldehyde groups are vulnerable to oxygen and sunlight (see, for example, Dunkelblum et al., 1984; Shani and Klug, 1980; Shaver and Ivie, 1982), and the attractivity of the pheromone is lost unless formulated with antioxidants and UV screens. (d)  Environmental Fate Using gas chromatography–mass spectrometry (GC–MS) detection, Spittler et al. (1992) could not detect residues on fruits exposed to aerial use dosages of mono- and diunsaturated alcohol acetate pheromones. In other experiments, however, pheromone absorption by the foliage could be demonstrated when the extremely sensitive natural “biosensor,” the male antenna, was used. For example, studying the absorption and release by apple leaves of a 95:5 mixture of (E)-11-tetradecenol acetate and (E,E)-9,11tetradecadienol acetate, the sex pheromone of the light brown apple moth (LBAM), Epiphyas postvittana, Karg et al. (1994) found that leaves could absorb enough of the pheromone blend from the airstream to serve as pheromone baits in the field. It has recently been found that sprayable formulations of pheromones and related aliphatic compounds, especially alcohols with 6–13 carbon atoms, cause local phytotoxicity presumably by nonspecific disruption of membranes of the treated plant leaves (Giroux and Miller, 2001). (e)  Biological Properties Mode of action  Insect pheromones are not true “insecticides” because they do not kill insects. They influence insect behavior through the olfactory system. In spite of decades of intensive electrophysiological and molecular biology studies, our understanding of the molecular determinants and mechanisms of pheromone perception and neural processing is limited (Blomquist and Vogt, 2003; Breer, 1997). Briefly, these volatile substances are adsorbed onto the surface of the antennae of the perceiving individual. The pheromone molecules then diffuse into the interior of the pheromone-tuned sensilla through

169

microscopic pores in the cuticle. Once inside, these lipophilic compounds are thought to be transferred through the aqueous sensillum lymph to the chemosensory membranes by pheromone-binding proteins (PBPs) with broad selectivity. Then the pheromone, or its PBP-complex, interacts with a pheromone-specific receptor protein, which then transduces the chemical signal into an electric signal. Sex pheromone perception at the molecular level is yet to be clarified, though evidence exists for the involvement of G-protein coupled receptors and/or ligand-gated cation channels. For multicomponent pheromone blends, specific PBPs, sensory neurons and receptors differentiate the components. The role of PBPs in the silk moth, Bombyx mori, has been well characterized (Gräter et al., 2006) and putative receptor proteins of a few lepidopteran species have been identified (see Mitsuno et al., 2008). For Drosophila, the situation appears to be more intricate since the maleproduced pheromone elicits a range of behavior both in females and males (reviewed by Benton, 2007). For continuous orientation, that is, upwind flight in a pheromone plume during mate searching, the insect also inactivates the pheromone and clears it from the perceiving antennae (reviewed by Vogt, 2005). Such rapid metabolic degradation was demonstrated for acetate (Ishida and Leal, 2005; Klun et al., 1996; Morse and Meighen, 1984; Vogt et al., 1985) and aldehyde (Morse and Meighen, 1984; Tasayco and Prestwich, 1990) lepidopteran pheromones, and a general kinetic model for the inactivation was also proposed (Kaisling, 1998). Toxicity and other biological effects  Acute toxicity data for pheromones and attractants are presented in Table 3.9. The data are representative but clearly indicate the low risk of the use of these substances. It should be mentioned that, compared to other insect control agents, only a handful of insect pheromones have been thoroughly examined for their toxic or other pharmacological effects on nontarget, including mammalian, species. Due to their low application rates, the expected low residues and low human exposure, pheromones and most other semiochemicals are considered low-risk pest control products. To ease the burden of costly toxicological evaluations regulatory agencies in many countries have developed special registration procedures, especially for straight chain lepidopteran pheromones for which the experience accumulated over the past decades has demonstrated safety. Field exposure in agricultural areas is expected to be limited but for occupational safety reasons each pheromone composition should preferably be evaluated on a caseby-case basis, especially for mating disruption formulations containing compounds with conjugated double bonds and/ or aldehyde or epoxide functionalities. Spray applications of pheromone formulations in residential areas are rare but require precautionary measures. Even in this situation human exposure is estimated to be low. A recent example

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Table 3.9  Acute Toxicity of Insect Pheromones and Attractants Compound

Oral LD50 (mg/kg), rat

Dermal LD50 (mg/kg), rabbit

Inhalation LC50 (mg/l), rat

References

(Z)-9-Tricosene

23,070

2025

26.6

Beroza et al. (1975)

(Z)-9-Tricosene

10,000

20,000

(Z)-9-Tricosene

5000

2000

5.0

U.S. EPA (1994)

()-Disparlure

34,600

2025

5.0

Beroza et al. (1975)

(Z)-7-Dodecenol

11,730

~3400

6.7

Beroza et al. (1975)

(E,E)-8,10-Dodecadienol

4000

5

Copping (2004)

(Z)-11-Hexadecenol

5600

(Z)-7-Dodecenol acetate

13,430

2025

Beroza et al. (1975)

(E)-9-Dodecenol acetate

15,000

3000

Copping (2004)

(E/Z)-4-Tridecenol acetate

5000

2000

(Z)-9-Tetradecenol acetate

5000

(Z)-7-Hexadecenol acetate

3460

(Z,E/Z,Z)-7,11-Hexadecadienol acetate

15,000

(Z,Z)-3,13-Octadecadienol acetate

5000

Inscoe and Ridgway (1992)

(Z)-9-Tetradecenal

5000

Inscoe and Ridgway (1992)

(E)-11-Tetradecenal

5000

(Z)-7-Hexadecenal

5000

Inscoe and Ridgway (1992)

(Z)-9-Hexadecenal

5000

Inscoe and Ridgway (1992)

(Z)-11-Hexadecenal

5000

(Z)-6-Heneicosen-11-one

15,000

3000

Trimedlure

4556

2025a

Hodosh et al. (1985)

Inscoe and Ridgway (1992)

5

Copping (2004) Inscoe and Ridgway (1992) Beroza et al. (1975)

2025 3.33

5000

16.9

5

Hodosh et al. (1985)

Hodosh et al. (1985)

Inscoe and Ridgway (1992) Hodosh et al. (1985)

2.9

Beroza et al. (1975)

a

One death out of four animals treated with 2025 mg/kg.

relating to the eradication program in California against the exotic insect LBAM, E. postvittana, serves as an illustration (California Office of Environmental Health Hazard Assessment, 2008; see also Werner et al., 2007). A microencapsulated slow-release formulation of the sex pheromone blend (see above) of this invasive insect was sprayed at total rates of about 200 g/ha at nights above infected areas, which included farms and residential areas with a potentially exposed population of about 392,000. After treatments, altogether 463 individuals reported respiratory symptoms alone or respiratory and dermal/ocular irritation. Analysis of the details and accounting for “background symptom reportings” showed that less than 10% of the reports could be related to exposure of the spray and even these were unlikely to be caused by the pheromone ingredient. Bedoukian (1992) tabulated the basic toxicological properties of fragrances and other volatile chemicals structurally related to insect pheromones. The toxicity

and nontarget effects of several lepidopteran sex pheromones (Inscoe and Ridgway, 1992) and coleopteran pheromones (Burke, 1992; see also Knipling, 1976) have been summarized. Ryan et al. (1992) tested 17 pheromone components for acetylcholine esterase inhibitory activity. The Ki values ranged from 0.27 mM for frontalin to 9.3 mM for ipsenol. Among the aliphatic sex pheromones, (Z)-11-hexadecenol acetate (Ki  0.37 mM) and (Z)-11-octadecenal (Ki  0.50 mM) were the most active. The gossyplure component (7Z,11Z)-7,11-hexadecadienol acetate had a Ki value of 1.75 mM. All compounds were reversible inhibitors of the enzyme. In an aquatic toxicity in vitro model that used the Gram-negative marine bacterium Vibrio fischeri, Cronin and Schultz (1998) found that the toxicity of pheromonelike, long chain aliphatic aldehydes and ketones correlated with their lipophilicity.

Chapter | 3  Pest Control Agents from Natural Products

Recently, the cytotoxicity of several insect sex pheromones, including long chain alkenals and alcohols as well as mono- and diunsaturated acetates and alcohols, to Chinese hamster ovary cells was evaluated in vitro (AbdelGhani et al., 2004; Bayoumi et al., 2002). All pheromones were cytotoxic with IC50 or equivalent inhibitory concentration values ranging from 22 to 178 M after 48 h incubation. It was also found that pre-exposure of the pheromones to rat liver submitochondrial fraction with oxidative properties attenuated the toxicity. Addition of bovine serum albumin or fetal calf serum in the culture medium also reduced the cytotoxicity of the compounds, indicating that protein binding that occurs in vivo reduces cellular bioavailability, thus cytotoxic side effects. Furthermore, the compounds were not mutagenic in the S. thyphimurium reversion test with or without submitochondrial pre-incubation. For products containing a combination of a pheromone or a semiochemical and an insecticide, the toxicity of the complete formulation should be considered. (f)  Specific Examples (Z)-9-Tricosene  IUPAC name: (Z)-tricos-9-ene. Chemical Abstract name: (9Z)-9-tricosene. CAS Registry Number: [27519-02-4]. Empirical formula: C23H46; molecular weight: 322.6. Pure muscalure is a colorless liquid with a boiling point of 157–158°C at 0.1 mmHg. The sex pheromone of the housefly, Musca domestica, was identified by Carlson et al. (1971) as monounsaturated hydrocarbon, (Z)-9-tricosene or muscalure (Figure 3.9). (Z)-9-Tricosene and related alkenes have also been detected in the cuticular hydrocarbon extract of other flies. Muscalure became the first pheromone to be registered as an insect control agent. Muscalure is used alone in baits or in combination with pyrethroid, carbamate or neonicotinoid insecticides (see, for example, Butler et al., 2007). The compound underwent extensive toxicity studies required for registration by the EPA (U.S. EPA, 1994). (Z)-9-Tricosene was practically nontoxic in a battery of avian acute and subacute tests. In reproductive toxicity studies with mallard ducks, 20 ppm of the 98.7% pure pheromone in the diet showed adverse effects for 3-week-old embryos, normal hatchlings, and 14-day-old survivors; the technical-grade material, however, produced abnormalities at 2 ppm. The pheromone was nontoxic to freshwater fish, but slightly toxic to Daphnia magna (LC50  1.08 ppm). The metabolism and environmental fate of (Z)-9-tricosene is less studied. In housefly cuticle, it undergoes CYPmediated oxidation forming ketone and epoxide metabolites (Ahmad et al., 1987). Disparlure  ()-Disparlure (racemic): IUPAC name: (7RS,8SR)-7,8-epoxy-2-methyloctadecane.

171

Chemical Abstract name: (2R,3S)-rel-2-decyl-3(5-methylhexyl)oxirane. CAS Registry Number: [29804-22-6]. ()-Disparlure: IUPAC name: (7R,8S)-7,8-epoxy-2-methyloctadecane. Chemical Abstract name: (2S,3R)-2-decyl-3(5-methylhexyl)oxirane. CAS Registry Number: [54910-51-9]. Empirical formula for both: C19H38O; molecular weight: 282.5. Females of the gypsy moth, Lymantria dispar, emit a powerful sex pheromone, attracting males of the same species from several hundred meters. The compound was identified, without establishing the absolute configuration, by Bierl et al. (1970). It was subsequently found that the ()-isomer (Figure 3.9) was responsible for the attractivity of the racemic synthetic mixture (Plimmer et al., 1977). In practice, however, the racemic mixture (henceforth disparlure) is used. Disparlure traps are used in forestry and in orchards for population monitoring and mass trapping while special slow-release formulations allowing the use of 15–100 g/ha were developed for mating disruption (Thorpe et al., 2006). The aerial disparlure concentration at these application rates was estimated as 3–30 ng/m3 (U.S. Department of Agriculture, 2006). Disparlure is essentially nontoxic and nonirritating (Table 3.9) (Beroza et al., 1975). In male antennae, the pheromone undergoes enzymatic hydrolysis, converting the epoxide into a behaviorally inactive diol (Prestwich et al., 1989). Cameron (1983) reported that years after being regularly exposed to synthetic disparlure, he became attractive to males of the gypsy moth, suggesting an unusual persistency of the lipophilic material in human skin. According to a toxicological risk assessment (U.S. Department of Agriculture, 2006), disparlure is not particularly toxic to fish and among the species tested rainbow trout was the most sensitive (72-h exposure to 100 ppm disparlure resulted in 20% mortality). The pheromone is slightly toxic to Daphnia magna (LC50  98 ppb). Gossyplure  IUPAC name: (7Z,11E)- and (7Z,11Z)-7,11hexadecadien-1-yl acetate. Chemical Abstract name: (7Z,11E)- and (7Z,11Z)-7,11hexadecadien-1-ol acetate. CAS Registry Numbers: unspecified stereochemistry: [50933–33–0]; (7Z,11E)-isomer: [53042-79-8]; (7Z,11Z)-isomer: [52207-99-5]. Empirical formula: C18H32O2; molecular weight: 280.4. The sex pheromone of the pink bollworm, Pectinophora gossypiella, a serious pest of cotton, was identified by Hummel et al. (1973) as a 1:1 mixture of (Z,E)- and (Z,Z)7,11-hexadecadien-1-yl acetates. It is now widely used in traps for monitoring as well as in special formulations

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172

for mating disruption alone or in combination with a neurotoxic insecticide (Haynes et al., 1986; Jackson, 1989; Jenkins, 2002). Henson (1977) found that the half-life of gossyplure in soil was 1 day, while in water it was 7 days. The loss was attributed to volatilization and hydrolysis to the corresponding alcohols.

Cl

O

2 4 5

O

1

(1S,2S,4R)-trans-trimedlure

(+)-α-copaene OH

Codlemone  IUPAC name: (8E,10E)-dodeca-8,10-dien-1-ol. Chemical Abstract name: (8E,10E)-8,10-dodecadien-1-ol. CAS Registry Number: [33956-49-9]. Empirical formula C12H22O; molecular weight: 182.3. Pure codlemone is a semisolid or colorless liquid with a melting point of 32°C; its boiling point is 110–120°C at 2 mmHg. Roelofs et al. (1971) identified this dienol (Figure 3.9) as a sex attractant of the codling moth, Cydia pomonella. The light-sensitive compound is extensively used in traps for population monitoring and also for mating disruption (reviewed by Witzgall et al., 2008).

3.2.3.2  Kairomones If the sex pheromone of a practically important insect species is not available or not suitable for practical purposes, another attractant semiochemical could offer an alternative solution. Two such compounds, a synthetic attractant and a plant-derived, nonvolatile phagostimulant, as representative kairomones will be mentioned. (a) Trimedlure Introduction  Intensive search for synthetic attractants of the Mediterranean fruit fly, Ceratitis capitata, resulted in the discovery of trimedlure (McGovern and Beroza, 1966). Trimedlure is a mixture of 16 stereo- and positional isomers of which the (1S,2S,4R)-4-chloro isomer (Figure 3.10) is the most attractive (see Warthen et al., 1995). The synthetic attractant is considered to be a structural analog of ()--copaene (Figure 3.10), a natural kairomone of the fly. Recently ceralure, an iodo analog of trimedlure, has emerged as a more effective and longer-lasting attractant (Jang et al., 2003). Identity  IUPAC name: tert-butyl ()-4(or 5)-chloro-2methylcyclohexanecarboxylate. Chemical Abstract: 4(or 5)-chloro-2-methylcyclohexanecarboxylic acid 1,1-dimethylethyl ester. CAS Registry Number: isomeric mixture: [12002-53-8]. Empirical formula: C12H21ClO2; molecular weight: 232.7. Trimedlure is a colorless volatile liquid with a fruity odor; its boiling point is 107–113°C. Toxicity to animals  In one of the first acute toxicity studies with attractants (Beroza et al., 1975), trimedlure

O

O H

HO 2

H

OH

O

1 R

O

O cucurbitacin A: 2β-OH, R = CH2OH cucurbitacin B: 2β-OH, R = CH3 cucurbitacin E: ∆1,2, R = CH3 Figure 3.10  Structures of natural and synthetic kairomones.

showed slight mammalian toxicity (Table 3.9). It caused local skin reactions characterized at the end of the 24-h contact period by erythema and edema. It was not an eye irritant in the Draize test. In static fish tests, bluegill sunfish became flaccid, with shallow respiration, and lay on the bottom of the tank. The 96-h LC50 was 12.1 ppm. With trout, trimedlure evoked dark discoloration of the integument, rapid and shallow respiration, excessive swimming with gyrating, and later lying on the bottom of the tank. The 96-h LC50 for this fish was 9.6 ppm. Formulations and uses  Traps containing up to 1–2 g of this attractant formulated in various dispensers are used for detection and mass trapping of C. capitata. Pharmacological studies  The 4-chloro-trans isomer of trimedlure effectively replaced the GABA-gated chloride channel probe [35S]t-butylbicyclophosphorothionate ([35S]TBPS) from receptors of housefly brain membrane preparations; rat brain [35S]TBPS receptors, however, were not sensitive to trimedlure (Cohen and Casida, 1985). (b) Cucurbitacins Introduction  Cucurbitacins are highly oxygenated, tetracyclic triterpenes present in the fruits and roots of cucurbits such as watermelon, squash, and zucchini at 0.1–0.3% concentrations (reviewed by Chen et al., 2005; Lavie and Glotter, 1971). These nonvolatile compounds are notable for their extreme bitterness with a detection level for humans of about 1 ppb in solution. They also have a broad range of pharmacological properties, including purgative, hepatoprotective, antifungal, anti-inflammatory, cytotoxic,

Chapter | 3  Pest Control Agents from Natural Products

and antineoplastic activities (reviewed by Miró, 1995). Importantly, they are locomotor arrestants and phagostimulants to Diabrotica species (Chrysomelidae), which are major coleopteran insect pests of maize in the United States, Mexico, and parts of Europe (Metcalf, 1994). Sprayable, cucurbitacin-containing insecticide baits, developed by Metcalf et al. (1987), greatly reduce insecticide application rates. Although it is obvious that the poisonous properties of this combination are due to the insecticide content, the plant material is also toxic and thus poses a risk during manufacture and handling. Identity  Of the dozens of cucurbitacins identified, representative examples for which relevant toxicological data exist are discussed only (Figure 3.10). Cucurbitacin A  Chemical Abstract name: (2,9,10,16, 23E)-25-(acetyloxy)-2,16,20-trihydroxy-9-(hydroxymethyl)19-norlanosta-5,23-diene-3,11,22-trione. CAS Registry Number: [6040-19-3]. Cucurbitacin B  Chemical Abstract name: (2,9,10,16, 23E)-25-(acetyloxy)-2,16,20-trihydroxy-9-methyl-19norlanosta-5,23-diene-3,11,22-trione. CAS Registry Number: [6199-67-3]. Cucurbitacin E  Chemical Abstract name: (9,10,16, 23E)-25-(acetyloxy)-2,16,20-trihydroxy-9-methyl-19-norlanosta-1,5,23-triene-3,11,22-trione. CAS Registry Number: [18444-66-1]. This compound (also called -elaterin) appears to be one of the most abundant in squash. Formulations and uses  The semiochemical–insecticide bait combination is formulated as either dry-flowable microspheres or polymer-based tank mixes. The source of the phagostimulant cucurbitacins used is usually the wild-growing buffalo gourd, Cucurbita foetidissima. Generally, commercial formulations contain about 13% carbaryl, the rest is plant material and inert ingredients (see, for example, Siegfried et al., 2004). Cucurbitacin-based baits containing other insecticides, including Bt toxins (Nowatzki et al., 2006) and phloxine B, a phototoxic xanthene dye (Schroder et al., 2001), have also been investigated. Acute toxicity to animals  The LD50 values upon intraperitoneal administration of cucurbitacin A were 1.2 mg/kg in male mice and 2.0 mg/kg in female rats; the LD50 value of cucurbitacin B was 1.0 mg/kg in mice (David and Vallance, 1955). Lethal doses caused respiratory distress and pathology showed acute pulmonary edema. Cucurbitacin E had an intraperitoneal LD50 of 2.0 mg/kg in mice (see Rymal et al., 1984). Stoewsand et al. (1985) reported diarrhea, anemia, and mortality in mice receiving a diet containing 1% cucurbita fruit of cultivars rich in cucurbitacin.

173

A multi-year study by Boetel et al. (2005) found that cucurbitacins-based insecticide formulations had no negative impact on nontarget arthropods. Human poisoning incidents  Ferguson et al. (1983), Kirschman and Suber (1989), and Rymal et al. (1984) summarized cucurbitacin-related food poisonings that occurred in the United States and Australia in the early 1980s. Cucurbitacins were considered to be responsible for uvular edema caused by the intranasal or oral applications of the juice of squirting cucumber, Ecballium elaterium, a popular folk remedy for treating inflammatory and other diseases (for recent case reports, see Caiozzi et al., 2002; Kavalci et al., 2007).

3.3  Disease Control Agents 3.3.1  Fungicides 3.3.1.1  Blasticidin-S (a)  Introduction Blasticidin-S is produced by the soil bacterium Streptomyces griseochromogenes (Actinomycetes). The structure of this peptidyl nucleoside antibiotic was elucidated by Õtake et al. (1966) (Figure 3.11) and its biosynthesis has also been clarified (Cone et al., 2003). Blasticidin-S was found to be fungicidal in 1958, and it has been used for the preventive and curative control of Pyricularia oryzae (teleomorph: Magnaporthe grisea), the causative agent of rice blast (reviewed by Yamaguchi, 1995, 1996). (b)  Identity, Physicochemical Properties, and Uses IUPAC name: 1-(4-amino-1,2-dihydro-2-oxopyrimidin-1yl)-4-[(S)-3-amino-5-(1-methylguanidino)valeramido]1,2,3,4-tetradeoxy--d-erythro-hex-2-enopyranuronic acid. Chemical Abstract name: (S)-4-[[3-amino-5-[(aminoimino methyl)methylamino]-1-oxopentyl]amino]-1-(4-amino2-oxo-1(2H)-pyrimidinyl]-1,2,3,4-tetradeoxy--derythro hex-2-enopyranuronic acid. CAS Registry Numbers: blasticidin-S: [2079-00-7]; blasticidin-S N-(4-benzylamino)benzenesulfonic acid salt: [51775-28-1]. Empirical formula: C17H26N8O5; molecular weight: 422.4. Physicochemical properties  Pure blasticidin-S forms colorless crystals melting at 253–255°C (with decomposition); the melting point of technical-grade material is 235–236°C (with decomposition). Blasticidin-S is dextrorotatory: [ ]11 D  108.4 (c  1.0 in water). Blasticidin-S is a weak base with pKa1  2.41 (carboxyl), pKa2  4.6, pKa3  8.0, and pKa4  12.5 (three bases). It is readily soluble in water and acetic acid (30 g/l in each at 20°C) but practically insoluble in common

Hayes’ Handbook of Pesticide Toxicology

174

NH2

NH2

HN

OH

N

O HO O

O

N

O

HN

H3C H2N O

HN

OH HO

NH2

N

HO

NH

O HO OH

O blasticidin-S

kasugamycin

OH

HN

NH2

OH

NH

N

HOOC

H2N

HO

H HO

N OH

O

O

OH

HO HO HN

HO

OH NH

H2N O

mildiomycin

HO

O

R

validamycin A R = β-D-Glc validoxylamine A R = H

Figure 3.11  Structures of microbial antifungal agents.

organic solvents. The pH of the aqueous solution of the free base is 9.3. The aqueous solution of the commercially available N-(4-benzylamino)benzenesulfonic acid (BABS) salt of the antibiotic has a pH of 6.0. Stability  Blasticidin-S is stable for 3 months at room temperature and at least 1 year at 4°C. It is stable in solution at a pH of 5–7, unstable at a pH less than 4, and decomposes under alkaline conditions with the loss of ammonia (Õtake et al., 1966). As dry film, blasticidin-S BABS salt is photostable. History, formulations, and uses  Blasticidin-S, the first fermentation-produced antibiotic developed for agricultural use, was isolated from the culture broth of S. griseochromogenes in 1955 and its unique fungicidal properties were discovered in 1958 (Takeuchi et al., 1958; reviewed by Misato, 1969; Yamaguchi, 1995). This nucleoside derivative possesses a wide range of biological activities, including antimicrobial (Takeuchi et al., 1958), antiviral (Hirai and Shimomura, 1965; Kummert and Semal, 1971), and antitumor (Tanaka et al., 1961) effects. Blasticidin-S at 10–40 g/ha application rates gives excellent control of the rice pathogen P. oryzae but its (phyto)toxic properties have lessened its significance.

The stable and nonphytotoxic blasticidin-S BABS salt is sold as dispersible powder, emulsifiable concentrate, or wettable powder formulations containing 1.4–6% active ingredient. To alleviate eye irritation (see following discussion), an improved formulation containing 5% calcium acetate additive was introduced (Yamaguchi, 1995). Blasticidin-S is also widely used in the laboratory to select transfected cells carrying resistance genes, which code for acetyl transferase or deaminase enzymes involved in the detoxification of the antibiotic. (c) Biological Properties Mode of action  Blasticidin-S inhibits protein biosynthesis in both prokaryotes and eukaryotes by interference with ribosomal peptidyl transfer (Kinoshita et al., 1970; Pestka et al., 1972). The recently solved x-ray crystal structure of blasticidin-S complexed with the large ribosomal subunit, known as 50S, of the bacterium Haloarcula marismortui revealed that the antibiotic binds at two overlapping binding sites of the peptidyl transferase (Hansen et al., 2003; Moore and Steitz, 2003). In a cell-free system of P. oryzae, the target pathogen, incorporation of amino acids into protein is inhibited, whereas other metabolic pathways, including glycolysis, electron transport, oxidative ­ phosphorylation,

Chapter | 3  Pest Control Agents from Natural Products

175

Table 3.10  Acute Toxicity of Blasticidin-Sa Species, sex

Assay

Rat

oral

39.5

Misato (1969)

Rat

oral

16.3

Yamashita et al. (1987)

Rat, male

oral

Rat, female

oral

Rat

dermal

Mouse

oral

Mouse Mouse, male

oral oral

LD50 (mg/kg)

Other data

References

b

Tomlin (2003)

b

55.9

Tomlin (2003)

500

Tomlin (2003)

56.8

10.1

Yamashita et al. (1987)

b

Yang and Deng (1996)

b

Tomlin (2003)

b

33.0

51.9

Mouse, female

oral

60.1

Tomlin (2003)

Mouse

iv

2.82

Takeuchi et al. (1958)

Carp

48-h

LC50  40 ppm

Copping (2004)

Daphnia pulex

3-h

LC50  40 ppm

Copping (2004)

a

Free base unless otherwise noted. For blasticidin-S N-(4-benzylamino)benzenesulfonate.

b

and nucleic acid synthesis, are not affected. Recent studies with the water mold, Achlya bisexualis, suggested inhibition of DNA synthesis as an additional effect (Sullia and Griffin, 1977). Degradation, metabolism, and excretion  On the plant surface, blasticidin-S is decomposed by sunlight and eventually gives rise to cytosine as the main degradation product (Yamaguchi et al., 1972). Common microbes in the field also contribute to the inactivation and disappearance of the antibiotic. An aminohydrolase (blasticidin-S deaminase, BSD), selectively catalyzing the deamination of cytosine, was isolated and characterized from resistant Aspergillus (Seto et al., 1966; Yamaguchi et al., 1975) and Bacillus (Endo et al., 1987) strains. Recently, Kumasaka et al. (2007) have reported the x-ray crystal structure of BSD engineered into E. coli. The residue level of blasticidin-S in rice was below 0.05 ppm 1 week after application, whereas the soil halflife of blasticidin-S was about 2 days under flooded conditions (Ebata, 1983). Upon oral application to the mouse, blasticidin-S and metabolites were excreted in the urine and feces within 24 h. (d)  Toxicity to Laboratory Animals Acute and chronic toxicity  In general, blasticidin-S is rather toxic to mammals but has low toxicity to fish. Acute toxicity data for several species are listed in Table 3.10. (Some reports fail to specify whether the material tested was the free base of blasticidin-S or its BABS salt, which could explain the variations in the reported LD50 values.)

In rats given blasticidin-S orally at 3 mg/kg or higher, alkaline phosphatase activity in serum and small intestine was temporarily reduced (see Ray, 1991). Pathology  The main pathological findings of blasticidin-S poisoning in animals relate to mucous membranes and the skin. Upon topical application, conjunctivitis, keratitis, nasal bleeding, and skin lesions, including hyperemia, edema and ulceration, are observed. Peritoneal adhesion involving intraabdominal organs and occasional gastrointestinal perforation are thought to be due to the lesions on the mucous membrane. Diarrhea is frequent and considered to be caused by irritation of the mucous membrane (Yamashita et al., 1987). Intratracheal injection of blasticidin-S into rabbits produced pneumonitis, characterized by focal destruction of tissues. Within 5–6 days, these proliferations formed glandular structures extending from the bronchiole. Within two more days, blood capillaries began to surround the glandular cells and then the glandular cells began to differentiate from the alveoli. By 12–14 days after treatment, both types of cells could be distinguished (Ebe, 1969). In a 2-year study with rats given up to 1 ppm blasticidin-S in the diet, no adverse effects were observed (Tomlin, 2003). The fungicide was nonmutagenic in a battery of S. typhimurium and E. coli bacterial reversion-assay systems (Moriya et al., 1983). (e)  Toxicity to Humans Irritation  Blasticidin-S causes irritation and inflammation upon contact with eyes and mucous membranes. A survey

176

of ophthalmic disturbances conducted in the 1960s showed that in certain years one-third of the applicators suffered some damage, including skin eruptions all over the body. Calcium acetate-containing formulations are safer in this respect (reviewed by Ray, 1991; Yamashita et al., 1987). Poisoning incidents  Yamashita et al. (1987) described four incidents of acute suicidal poisoning from ingested blasticidin-S. In the three fatal cases, nausea, vomiting, and severe diarrhea appeared almost immediately after ingesting 100–250 ml of undiluted formulations containing 2–5 g blasticidin-S BABS salt. Pain in the oral cavity and pharynx was noted in all patients. In one case, the vomitus was bloody and esophageal pain was claimed. All patients were completely conscious and restless during the 6–10 h after ingestion. Hypotension associated with tachycardia became more pronounced as time passed. Marked cold, pale, and perspired extremities were usually noted. Thermal symptoms indicated insufficient peripheral circulation. No cardiac anomalies were seen. The average hematocrit and hemoglobin concentrations for the three fatal cases were 52.9 and 17.4 g/100 ml, respectively. Remarkable hemoconcentration was consistently noticed. Laboratory findings revealed moderate hepatic dysfunction. In the terminal phase of the fatal cases, blood pressure dropped and pulse rates increased above 120 beats/min. Death occurred about 1 day after ingestion. Toyoshima et al. (1994) reported an unusual case of acute interstitial pneumonia caused by inhalation of blasticidin-S powder, while Tamura et al. (2000) described a poisoning case in which coagulopathy was observed. Yang and Deng (1996) gave a detailed analysis of 24 suicidal, three occupational and one accidental poisoning cases that occurred in Taiwan between 1985 and 1993. None of the five fatalities was work related. The characteristic symptoms of blasticidin-S poisoning again were gastrointestinal disorders; redness of the conjunctiva; hypotension, occasionally preceded by hypertension, with tachy- or bradycardia; and aspiration pneumonia. Occasionally, neurological manifestations of poisoning could be seen. As little as 70 mg of the active ingredient was capable of producing symptoms. Death resulted from cardiovascular collapse or aspiration pneumonitis with possible bronchospasm. Treatment  Because blasticidin-S is water soluble, contaminated skin should be washed. The oral and nasal cavities should be cleaned as well. Prompt symptomatic treatment is important and should include intravenous fluid administration and management of water and electrolyte balance. Hydration and adequate urinary output are essential to facilitate renal elimination of the toxicant. Direct hemoperfusion and plasma exchange has also been effective (Tamura et al., 2000). Prevention of aspiration pneumonitis and support of ventilatory function are also vital. For workers with chronic exposure, steroids and antibiotics could be helpful.

Hayes’ Handbook of Pesticide Toxicology

3.3.1.2  Kasugamycin (a)  Introduction Kasugamycin (Figure 3.11) is an aminoglycoside antibiotic produced by Streptomyces kasugaensis (Umezawa et al., 1965). Its structure was established by Suhara et al. (1966). Kasugamycin has systemic activity and is widely used to control the rice blast disease caused by P. oryzae (reviewed by Yamaguchi, 1995, 1996). (b)  Identity, Physicochemical Properties, and Uses IUPAC name: 1L-1,3,4/2,5,6-1-deoxy-2,3,4,5,6-pentahydroxycyclohexyl 2-amino-2,3,4,6-tetradeoxy-4-(-iminoglycino)--D-arabino-hexopyranoside. Chemical Abstract name: 3-O-[2-amino-4-[(carboxyimino methyl)amino]-2,3,4,6-tetradeoxy--D-arabino-hexopyranosyl]-D-chiro-inositol. CAS Registry Number: kasugamycin [6980-18-3]; kasugamycin hydrochloride hydrate: [19408-46-9]. Empirical formula: kasugamycin: C14H25N3O9; molecular weight: 379.4; kasugamycin hydrochloride hydrate: C14H28ClN3O10; molecular weight: 433.8. Physicochemical properties  The HCl hydrate forms sweet, colorless crystals melting at 202–204°C (with decomposition). At room temperature, the solubility of kasugamycin HCl hydrate is 220 g/l in water and 7.4 mg/l in methanol; it is insoluble in acetone, ethyl acetate, and chloroform. Kasugamycin is a weak base with pKa1  2 (carboxyl), pKa2  7.1, and pKa3  10.6 (two bases). Kasugamycin is dextrorotatory: [ ]25 (c  1.6 in water). D  120 Stability  Kasugamycin HCl hydrate is more stable than the free base and does not deteriorate upon storage at 50°C for 10 days. It tolerates weak acids, decomposes slowly at pH  7, but decomposes within weeks in alkaline solutions even at ambient temperature. Formulations and uses  Kasugamycin is produced by fermentation and usually sold as hydrochloride hydrate. Typical formulations for spraying, dusting, or seed treatment are wettable powders and soluble liquid concentrates alone (0.3–3% active ingredient) or in combination with other pesticides, such as copper oxychloride (Copping, 2004; U.S. EPA, 2005c). In addition to controlling rice blast and a few other fungal diseases of potato, pepper and tomato, kasugamycin is active against Pseudomonas, Erwinia, Xanthomonas, and Corynebacterium bacterial species (Ogawa, 1992). (c)  Biological Properties Mode of action  In cell-free systems, kasugamycin inhibited protein synthesis in P. oryzae and Pseudomonas fluorescens markedly but much less so in rat liver preparation (Tanaka et al., 1965). The antibiotic was shown to interfere with ­

Chapter | 3  Pest Control Agents from Natural Products

peptidyl–tRNA binding to the 30S ribosomal subunit of E. coli resulting in the inhibition of translation initiation (Okuyama et al., 1975). Recent X-ray crystallography studies have revealed two antibiotic binding locations: one at the peptidyl–tRNA site and another at the exit–tRNA site (Schluenzen et al., 2006; Schuwirth et al., 2006). Interestingly, kasugamycin is inhibitory to P. oryzae in acidic (pH  5) but not in neutral media (Hamada et al., 1965). Metabolism and excretion  Oral administration of 100 mg/kg kasugamycin to mice indicated rapid absorption and 43–68% excretion with the urine in 6 h (Takeuchi et al., 1965). When a rabbit was subcutaneously injected with the same dose of kasugamycin, the fungicide disappeared from the blood within 8 h, and 96% of the injected material was excreted with urine within 8 h after injection; kasugamycin concentration in the urine was highest (43 mg/ml) after 45 min. On intramuscular injection of 1.0 g kasugamycin into humans, about 63% of the fungicide was excreted unchanged with urine within 8 h. In rats, 90% of the orally administered fungicide was excreted with the feces and 2–3% with the urine by 168 h (U.S. EPA, 2005c). In plants, such as tomato, the major metabolic pathway of fungicide involved conjugation and hydrolytic metabolism, though the major plant residue is the intact fungicide. (d)  Toxicity to Laboratory Animals Acute and chronic toxicity  Acute toxicity data for kasugamycin HCl hydrate are shown in Table 3.11. Doses of 2000 mg/kg kasugamycin, administered intravenously, subcutaneously, or intraperitoneally to mice, caused neither observable effects nor death (Takeuchi et al., 1965; see also

177

Matsuzaki et al., 1968). Furthermore, an 800 mg/kg dose intravenously injected into a monkey failed to show any toxic effects. The blood analysis of monkeys receiving repeated intramuscular doses up to 800 mg/kg of the fungicide was normal. Similarly, no effect on the blood chemistry was found upon administration of a total of 10 g kasugamycin to a dog (10 kg) during a 45-day trial. Kasugamycin hydrochloride hydrate was found to be a mild eye irritant in the rabbit study. Anal lesions and perigenital staining were observed in mouse subchronic studies (U.S. EPA, 2005c). These effects are probably due to the acidic character of the fungicide. Kasugamycin, belonging to the aminoglycoside antibiotics known for their potential nephrotoxic, ototoxic, and neuromuscular paralytic activity, did not induce polymerization of rabbit muscle actin in vitro (Someya and Tanaka, 1979). Furthermore, studies in vitro indicated ototoxicity for streptomycin but not for kasugamycin (Masuko et al., 1999). In chronic dietary toxicity studies in rats, only the highest (3000 ppm) dose indicated adverse effects, namely testicular softening and atrophy in males, although the increased incidences were not statistically significant; the NOEL is 11.3 mg/kg daily dose of the fungicide for both males and females (U.S. EPA, 2005c). Kasugamycin was nonmutagenic in a battery of S. typhimurium and E. coli bacterial reversion-assay systems and no carcinogenic effects were found (Moriya et al., 1983; U.S. EPA, 2005c). (e)  Toxicity to Humans Based on the remarkable low toxicity of kasugamycin, it was tested and proven to be effective against Pseudomonas aeruginosa urinary infections in humans (Takeuchi et al., 1965).

Table 3.11  Acute Toxicity of Kasugamycin Hydrochloride Hydrate Species, sex

Assay

LD50 (mg/kg)

Rat

oral

5000

Ogawa (1992)

Rat

dermal

5000

Ogawa (1992)

Rat

4-h inhalation

LC50  2.4 mg/l air

Ogawa (1992)

Rat

4-h inhalation

LC50  4.9 mg/l air

U.S. EPA (2005c)

Rabbit

dermal

2000

Copping (2004)

Monkey

iv

800

Takeuchi et al. (1965)

Japanese quail

oral

4000

Copping (2004)

Carp

48-h

LC50  40 ppm

Copping (2004)

Goldfish

48-h

LC50  40 ppm

Copping (2004)

Daphnia pulex

6-h

LC50  40 ppm

Copping (2004)

LD50  40 g/bee

Copping (2004)

Honeybee

Other data

References

178

3.3.1.3  Mildiomycin (a)  Introduction The antibiotic mildiomycin was isolated from the culture broth of the actinomycete Streptoverticillium rimofaciens (Iwasa et al., 1978). Mildiomycin is a peptidyl nucleoside derivative with an arginine-like side chain and a serine residue connected to 5-hydroxymethylcytosine (Harada and Kishi, 1978; Harada et al., 1978) (Figure 3.11). It is a potent and selective fungicide used against various pathogens causing powdery mildews in fruits and vegetables (Iwasa, 1983; Kusaka et al., 1979). The biosynthesis of the antibiotic has been studied (see, for example, Li et al., 2008). (b)  Identity, Physicochemical Properties, and Uses IUPAC name: (2R,4R)-2-{(2S,3S,6R)-6-[4-amino-1,2-dihydro-5-(hydroxymethyl)-2-oxopyrimidin-1-yl]-3,6-dihydro4-(L-serylamino)-2H-pyran-2-yl}-2,4-dihydroxy-5-guanidinopentanoic acid. Chemical Abstract name: (S)-4-amino-1-[4-[(2-amino-3hydroxy-1-oxopropyl)amino]-9-[(aminoiminomethyl) amino]-6-C-carboxy-2,3,4,7,9-pentadeoxy--L-talo-non2-enopyranosyl]-5-(hydroxymethyl)-2(1H)-pyrimidinone. Code numbers: Antibiotic B-98891, TF-138. CAS Registry Number: [67527-71-3]. Empirical formula: C19H30N8O9; molecular weight: 514.5. Physicochemical properties  The melting point of mildiomycin hydrate is greater than 300°C (with decomposition). Mildiomycin is hygroscopic, readily soluble in water and acids, but sparingly soluble in dioxane, dimethyl sulfoxide, and pyridine. Mildiomycin is a weak base with pKa1  2.8 (carboxyl), pKa2  4.2, pKa3  7.2, and pKa4  12 (three bases). The antibiotic is dextrorotatory: [ ]20 D  100 (c  0.5 in water); (c  0.5 in 0.1 N HCl). []20   78 . 5  D Stability  In aqueous solutions, mildiomycin is stable at pH  7, but slowly decomposes in alkaline (pH  9) and strongly acidic (pH  2) media. Formulations and uses  The fungicide is produced by fermentation and formulated as a wettable powder or aqueous solution containing 8% active ingredient. It was introduced against powdery mildews in cucumber, apple, grape, barley, green pepper, strawberry, mulberry, tobacco, and rose. It is applied to the foliage as a spray at concentrations of 40–80 ppm (Kusaka et al., 1979). (c)  Biological Properties Mode of action  Mildiomycin suppressed the growth of E. coli by inhibiting bacterial polypeptide synthesis without interfering with respiration, oxidative phosphorylation, nucleic acid synthesis, or lipid and steroid biosynthesis. Polypeptide synthesis in a mammalian cell-free system

Hayes’ Handbook of Pesticide Toxicology

from rabbit reticulocytes was less sensitive than the bacterial system from E. coli (Om et al., 1984). Mildiomycin selectively inhibits protein synthesis by blocking the peptidyltransfer in human HeLa cells. RNA or DNA synthesis is not affected. When HeLa cells are permeabilized by animal viruses, the inhibitory effect on protein synthesis increases, indicating that the basis for the selective antibacterial action of mildiomycin could be due, at least in part, to poor penetration into the cell (Feduchi et al., 1985). (d)  Toxicity to Laboratory Animals Acute and chronic toxicity  Although mildiomycin is a structural relative of blasticidin-S, the compounds differ greatly in their toxicities to plants and mammals. Mildiomycin has a very low acute toxicity to test animals (Table 3.12). In 30-day feeding trials, no treatment-related adverse effects were observed in mice or rats at 200 mg/kg daily doses. In a 3-month subacute study, the daily NOEL was 50 mg/kg in rats. The antibiotic was nonmutagenic in the Ames test with or without rat liver homogenate (Kusaka et al., 1979).

3.3.1.4  Validamycin A (a)  Introduction The pseudotrisaccharide validamycin A (Figure 3.11) is the major antifungal component of the validamycin complex isolated from the culture broth of Streptomyces hygroscopicus subsp. limoneus (Iwasa et al., 1971a). The first proposed structure of this aminosugar antibiotic (Horii and Kameda, 1972) was later revised by Suami et al. (1980). In China, the same antibiotic, isolated from S. hygroscopicus subsp. jinggangensis, is named jinggangmycin. The biosynthesis of validamycin and related compounds has been studied extensively (see, for example, Bai et al., 2007). (b)  Identity, Physicochemical Properties, and Uses IUPAC name: (1R,2R,3S,4S,6R)-2,3-dihydroxy-6-hydroxymethyl-4-[(1S,4R,5S,6S)-4,5,6-trihydroxy-3hydroxymethylcyclohex-2-enylamino]cyclohexyl -d-glucopyranoside. Chemical Abstract name: 1,5,6-trideoxy-4-O--d-glucopyranosyl-5-(hydroxymethyl)-1-[[(1S,4R,5S,6S)-4,5,6trihydroxy-3-(hydroxymethyl)-2-cyclohexen-1-yl] amino]-d-chiro-inositol. CAS Registry Number: [37248-47-8]. Empirical formula: C20H35NO13; molecular weight: 497.5. Physicochemical properties  Validamycin A is a colorless hydrophilic powder without a sharp melting point: It softens at approximately 100°C and decomposes at approximately 135°C. The antibiotic is readily soluble in water, methanol, and dimethyl sulfoxide; sparingly soluble in ethanol

Chapter | 3  Pest Control Agents from Natural Products

Table 3.12  Acute Toxicity of Mildiomycina Species, sex

Assay

LD50 (mg/kg) Other

Rat, male/female

oral

4300/4120

Rat, male/female

sc

463/684

Rat, male/female

iv

885/700

Rat, male/female

ip

679/842

Rat

dermal

5000

Mouse, male/female

oral

5060/5250

Mouse, male/female

sc

1190/1150

Mouse, male/female

iv

645/599

Mouse, male

ip

1020/1050

Mouse

dermal

5000

Carp

72-h

LC50  40 mg/l

Japanese killifish

7-day

LC50  40 mg/l

Daphnia pulex

6-h

LC50  20 mg/l

a

From Kusaka et al. (1979).

and acetone; and insoluble in ethyl acetate and ethyl ether. Validamycin A is dextrorotatory: [ ]24 D  110  15 (c  1 in water). Its pKa is 6.0. Validamycin A monohydrochloride is a colorless crystalline powder with a melting point of 95°C (with decomposition). The salt is soluble in water, methanol, and dimethyl sulfoxide, slightly soluble in acetone and ethanol, and insoluble in ethyl acetate and ethyl ether. The optical rotation of the salt is [ ]22 D  49  10 (c  1 in water). Stability  Validamycin A is stable in mild alkaline and acidic solutions. It is stable in sunlight, but in soil microbial degradation is rapid with a half-life of less than 2 h (Asano et al., 1984; Matsuura, 1983). Formulations and uses  Validamycin A is produced by fermentation and formulated as 3% liquid concentrate and 0.3% dust in foliar sprays, or for soil incorporation and seed dressing. It is widely used in Asia for the treatment of sheath blight of rice, black scurf on potatoes, bottom rot on lettuce, and against other diseases caused by Rhizoctonia solani and other basidiomycetes fungi. Enzymatic decomposition of validomycin affords the simple -d-glucose-like unsaturated aminocyclitol valienamine, an important precursor of the antidiabetic drug, voglibose. (c)  Biological Properties Mode of action and biochemical effects  The mode of action of validamycin A was reviewed by Yamaguchi (1995). The antibiotic alters the morphology of R. solani

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by inhibiting the biosynthesis of myo-inositol, thereby reducing the pathogenicity of the fungus. Validamycin A was found to inhibit trehalases from R. solani (Asano et al., 1987) as well as from various other organisms, including rat, rabbit, pig, yeast and insect with IC50 values ranging from 108 to 106 M (Kameda et al., 1987). Trehalase (,-trehalose glucohydroxylase, EC 3.2.1.28) is widespread among many organisms and is important in regulating d-glucose (energy) supply for flight muscle in insects, for germination of spores, etc. In mammals, the enzyme is found in intestinal and kidney membranes and its function is to hydrolyze ingested trehalose. Validamycin A is a poor inhibitor (IC50  103 M) of other sugar-hydrolyzing enzymes such as porcine intestinal maltase, isomaltase, and sucrase (Kameda et al., 1986). In R. solani, the antibiotic was shown to be efficiently transported into the mycelia and hydrolyzed therein by a -glycosidase to validoxylamine A [38665-10-0], a more potent inhibitor of trehalase of the fungus both in vitro and in vivo (Asano et al., 1987). The recently solved x-ray crystal structure of the complex of trehalase with validoxylamine A revealed that this tight-binding inhibitor mimics the transition state of the disaccharide trehalose at the active site of the glycosidase (Gibson et al., 2007). Validamycin A lacked fungicidal activity in vitro against Fusarium oxysporum f. sp. lycopersici but by foliar application it controlled tomato wilt, which is caused by this soil-borne pathogen, indicating again the requirement of bioactivation (Ishikawa et al., 2005). However, another mode of action for the antibiotic has recently been proposed. Since validamycin A is not a systemic compound, Ishikawa et al. (2007) suggest that the lasting effect of fungicide is due to the activation of the systemic acquired resistance (SAR) of the plant as indicated by necrosis of the affected tissues and elevations of SAR markers, such as salicylic acid. Validoxylamine A strongly inhibits mammalian intestinal, yeast, and insect trehalases also (Kameda et al., 1987; Kyosseva et al., 1995) and affects the activity of catalase, acid phosphatase and urease of soil microorganisms as well (see Qian et al., 2007). Certain sugars, such as l-sorbose (Trinci, 1985), fructose, glucose, sucrose, lactose, and mannose (Robson et al., 1991), antagonized the growth inhibitory effect of validamycin A in Rhizoctonia species. Metabolism  Rats upon oral administration, rice plant, bacteria, and soil metabolize validamycin A into validoxylamine A and D-glucose. Rats on intravenous administration, however, excrete the intact antibiotic in the urine (Kameda et al., 1975; Matsuura, 1983). (d)  Toxicity to Laboratory Animals Acute and chronic toxicity  Validamycin A has low toxicity to mammals. The oral and subcutaneous LD50 values

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for rats and mice are greater than 20,000 and 15,000 mg/kg, respectively. The intraperitoneal and intravenous LD50 values in mice are greater than 13,000 and 10,000 mg/kg, respectively. No irritative effects on the skin at 10 mg/cm2 and on the cornea at 10 mg/eye of rabbits were observed. Acute exposure of rats to 12.46 mg/l air of validamycin A aerosol caused no untoward reactions during exposure for 14 days afterwards. Oral administration of 12.5 g/kg doses to chicken and quail showed no treatment-related effects. The antibiotic at 10,000 ppm in the diet did not show any effect in rats and mice in 23-day subchronic studies. In a 2-year feeding study with rats, no treatment-related effects were seen for 1000 ppm (40.4 mg/kg) daily doses. Validamycin A at 40 and 10,000 ppm did not affect carp and killifish, respectively, in a 72-h toxicity assay (Anon., no date; Matsuura, 1983; Tomlin, 2003; see also Iwasa et al., 1971b). Onishi and Miyaji (1973) summarized the results of 3month feeding experiments with rats and mice given validamycin A in food at 0.1–10% concentrations. For both species, the highest dose evoked an increased tendency of diarrhea from the third day, lasting for about 2 months. Analyses of the blood and urine of rats showed only minor treatment-related changes, generally in males. Pathological examinations revealed hypermucosecretion of the cecum, pneumonia, focal granulomatous formations in cardial tissues, and hepatic congestions for both rodent species, but these abnormalities were slight and sporadic. Reproduction, teratology, and mutagenicity studies  When rats were fed 500 and 10,000 ppm validamycin A, no deaths or abnormalities were seen in the F0 and F1 generations or in teratology studies of the F1(b) and F2(b) progeny. The fungicide was nonmutagenic in a battery of S. typhimurium and E. coli bacterial reversion-assay systems (Moriya et al., 1983).

3.3.2  Bactericides 3.3.2.1  Streptomycin (a)  Introduction Streptomycin (Figure 3.12) was discovered as a fermentation product of the soil actinomycete Streptomyces griseus in 1943 (see Waksman, 1953). The aminoglycoside structure of the antibiotic was subsequently established (Kuehl et al., 1948; Wolfrom et al., 1954). Once a widely used antibacterial agent both in medicine and plant disease control, streptomycin has lost its importance due to toxicity and resistance problems. The derived dihydrostreptomycin is also commercialized. Streptomycin is a well-studied antibiotic and is discussed, among others, in standard textbooks (see, for example, Chambers, 2006). This overview will concentrate on recent works pertinent to the agricultural use and general toxicology of this antibiotic.

NH2

HN NH2

NH

HN N H O R

HO

OH

OH

O

H3C OH

O

HO O HO

HO

streptomycin

NHCH3

R = CHO

dihydrostreptomycin R = CH2OH Figure 3.12  Structures of streptomycin and its reduced derivative.

(b)  Identity, Physicochemical Properties, and Uses IUPAC name: O-2-deoxy-2-methylamino--l-glucopyranosyl-(1→2)-O-5-deoxy-3-C-formyl--l-lyxofuranosyl-(1→4)-N1,N3-diamidino-d-streptamine. Chemical Abstract name: O-2-deoxy-2-(methylamino)-l-glucopyranosyl-(1→2)-O-5-deoxy-3-Cformyl--l-lyxofuranosyl-(1→4)-N,N-bis (aminoiminomethyl)-d-streptamine. CAS Registry Numbers: streptomycin [57-92-1]; streptomycin sesquisulfate [3810-74-0]; dihydrostreptomycin [128-46-1]. Empirical formulas: streptomycin: C21H39N7O12, molecular weight: 581.6; streptomycin sesquisulfate: C42H84N14O36S3, molecular weight: 1457.3; dihydrostreptomycin: C21H41N7O12, molecular weight: 583.6. Physicochemical properties  The antibiotic is usually available as its sesquisulfate salt, under the name of “streptomycin sulfate,” which is a white to light-gray, hygroscopic powder with a faint amine-like odor. The solubility of the sulfate salt in water is greater than 20 g/l; in ethanol, the solubility is 0.35 g/l, and it is essentially insoluble in ethyl ether. Streptomycin sesquisulfate is levorotatory: []D  79.5 (c  1% in water). Stability  At ambient temperature, streptomycin is stable between pH  3 and pH  7 but degrades in strong acids and alkalis. Its salts are relatively stable to heat or light although deliquesce on air. Accordingly, streptomycin residues in honey were found to be stable for several months (Pang et al., 2004). Formulations and uses  Streptomycin sesquisulfate is formulated as a wettable powder and sprayable liquid with

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1–62% active ingredient alone or in combination with another antibacterial or fungicidal agents. Streptomycin, one of the few antibiotics used against plant pathogens, is active against aerobic Gram-negative bacteria. Although resistance, phytotoxicity, and restrictive legislation severely limit the use of streptomycin, it is still applied to control fire blight (causative agent Erwinia amylovora) on apples and pears and various other bacterial diseases in rice, nurseries, stone fruits, tobacco, some vegetables and ornamentals at rates ranging from 50 to 2000 ppm (30–500 g/ha) (for a review, see McManus et al., 2002). Streptomycin may be mixed with iron chloride or citrate to reduce its phytotoxicity. To combat resistance, it is often applied in combination with other bactericides such as oxytetracycline. The EPA established tolerances or maximum residue limits of 0.25 ppm for residues of streptomycin in or on vegetables and pome fruits (U.S. EPA, 2006). (c)  Biological Properties Mode of action  The principal mode of antibacterial action of streptomycin is the inhibition of protein synthesis at the aminoacyl-transfer site of the 16S portion of the 30S ribosomal RNA subunit that results in misreading of the genetic code. This was corroborated by recent crystal structure analyses (Carter et al., 2000). Additional sites of binding to ribosomes have also been described (Spickler et al., 1997). These and other mechanisms, such as membrane damage, involved in the antibacterial effect of streptomycin were reviewed by Chambers (2006) and Kornder (2002). Metabolism and excretion  The metabolism and excretion of streptomycin and related aminoglycosides in animals were summarized by Huber (1988). Streptomycin is poorly absorbed upon ingestion and excluded from most tissues, including the CNS. The drug is not metabolized in the intestine and at least two-thirds of it can be recovered in the urine and feces. The antibiotic is absorbed more readily upon intramuscular injection, reaching a peak concentration in blood within 1 h. About two-thirds of the intramuscularly administered streptomycin is excreted via urine within 24 h, but small amounts can be detected in the urine even 10–20 days after injection. Accumulation of streptomycin and the other aminoglycoside antibiotics can be observed in the kidney and the inner ear. In an exotic, comparative study, the peak serum concentrations of streptomycin following intramuscular injection in camel, cattle, pig, and dog were 30, 60, 60, and 60 min, respectively. In camels, for an intravenous dose of 10 mg/ kg, the serum half-life was about 200 min and the systemic clearance was 0.93 ml/min/kg (Hadi et al., 1998). (d)  Toxicity to Laboratory Animals and Wildlife Acute and chronic toxicity  The acute oral LD50 of streptomycin for mice is greater than 10,000 mg/kg; the acute

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intraperitoneal LD50 values for male and female mice are 340 and 305 mg/kg, respectively; and the acute dermal LD50 values for male and female mice are 400 and 325 mg/ kg, respectively. The acute oral LD50 values of streptomycin sesquisulfate for mice and rats are 9000 mg/kg for each (Tomlin, 2003). Doretto et al. (1994) reported rat intraperitoneal LD50 values of 1219 mg/kg for streptomycin sesquisulfate and 2023 mg/kg for streptomycin hydrochloride–calcium chloride double salt. Streptomycin is practically nontoxic to birds and honeybees, but slightly toxic to fish with a 96-h LC50 value of 180 ppm for rainbow trout (U.S. EPA, 1992). In doses above 70 mg/kg, streptomycin can produce a paralytic neuromuscular blockade in cats and dogs, similar to the effect of magnesium (see Rance and Randall, 1986). At 220–440 mg/kg intravenous doses, streptomycin causes irreversible lowering of systemic arterial pressure, probably via depression of the vasomotor center. The symptoms of acute streptomycin poisoning include restlessness, nausea, labored respiration, loss of consciousness, and coma (Huber, 1988). The streptomycin hydrochloride–calcium chloride complex was found to be an equipotent and safer substitute for streptomycin sulfate as determined by nystagmus changes to rabbits upon chronic administration (Doretto et al., 1994). Based on chronic dietary and reproductive toxicity animal studies carried out in connection with the pharmaceutical use of streptomycin, a daily NOEL value of 0.05 mg/kg chronic dietary dose for humans was established (U.S. EPA, 2006; see also Lemeire et al., 2007). The use of streptomycin in crop protection is limited as is the exposure of nontarget organism to the antibiotic. Ecotoxicological data, especially those concerning aquatic ecosystems, originate mainly from studies related to its use in veterinary medicine (see, for example, Taub et al., 1983). Thus, for Daphnia magna the 48-h LC50 is 487 ppm (Wollenberger et al., 2000), while for the freshwater green alga, Selenastrum capricornutum, and the freshwater cyanobacterium, Microcystis aeruginosa, the EC50 values are 0.133 and 0.007 ppm, respectively (Halling-Sørensen, 2000). Biochemical effects and pharmacology  Disruption of ribosomal functions is but one of a multitude of biological activities demonstrated for aminoglycoside antibiotics in vitro. Aminoglycosides, in general, interfere with various phospholipases, sphingomyelinases, ATPases, and intracellular second messengers (Chambers, 2006). Streptomycin, by virtue of its basic guanidine moieties, may also perturb the structure of cell membranes, for example, by interacting with anionic phospholipids. Wagner et al. (1987) demonstrated that the neuromuscular toxicity of streptomycin and other aminoglycoside antibiotics correlated with the inhibition of Ca2 uptake by specific voltage-gated ion channels.

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Studies using the phrenic nerve isolated from the rat showed that aminoglycosides produce neuromuscular blockade that could not by reversed by neostigmine but was revoked by calcium. In electrophysiological experiments, streptomycin reversibly blocked ACh-induced current in cochlear hair cells in guinea pigs, possibly by blocking the Ca2 entry necessary to elicit cholinergic responses (for a discussion, see Lima da Costa et al., 1998). Furthermore, Rothlin et al. (2000) have recently reported that streptomycin as well as other ototoxic aminoglycosides block the 9 subunit-containing recombinant rat nACh receptor in vitro in a noncompetitive manner with regard to ACh and this blockade was again reversible with increasing extracellular Ca2 concentration. It was also postulated that the cochleotoxicity and vestibulotoxicity produced by aminoglycoside antibiotics involve the excitotoxic activation of the cochlear N-methyld-aspartate (NMDA) subtype of glutamate receptors (Basile et al., 1999; Masuko et al., 1999). Streptomycin inhibited voltage-gated Ca2 channels in various test systems (Miller and Langton, 1998, and references therein) and had anti-arrhythmic effects in the rat heart (Salmon et al., 1997). Using the polymerization of rabbit skeletal muscle actin as a model for neuromuscular paralysis, Someya and Tanaka (1979) showed that streptomycin could induce polymer formation of actin and suggested that a related mechanism could be responsible for the neuromuscular side effect of the antibiotic. Recently, streptidine, the diguanidinocyclitol metabolite of the antibiotic formed in vulnerable individuals, has been implicated in the ototoxicity of streptomycin (see Granados and Meza, 2007). In rats, urinary -glutamyl transferase activity was shown to be a useful indicator of acute streptomycin-induced nephrotoxicity (Dierickx, 1981). Treatment  The treatment of choice for acute streptomycin poisoning is artificial respiration coupled with intravenous administration of calcium chloride, as demonstrated in cats (Sarkar et al., 1992). Calcium salts of pantothenic acid derivatives increased the therapeutic ratio by markedly reducing the acute toxicity of streptomycin in rats and mice (Dorofeev et al., 1987; see also Kornder, 2002). The common protein synthesis inhibitor cycloheximide was recently shown to prevent or reduce streptomycininduced apoptosis and DNA damage of vestibular hair cells (Nakagawa et al., 1998). Two NMDA antagonists, the voltage-dependent channel blocker dizocilpine and the polyamine antagonist ifenprodil, were also found to attenuate the aminoglycosideinduced hearing loss and the destruction of cochlear hair cells in guinea pigs (Basile et al., 1996). (e)  Toxicity to Humans Over the past five decades, a considerable amount of experience has been accumulated on the untoward side

Hayes’ Handbook of Pesticide Toxicology

effects of streptomycin use as an antitubercular agent (Bagger-Sjöbäck, 1997; Chambers, 2006; Norris, 1988). Streptomycin, dihydrostreptomycin, and other aminoglycoside antibiotics are notorious for their ototoxicity and nephrotoxicity, which can be irreversible at chronic doses. The acute effects, including neuromuscular blockade and hearing impairment, are mostly reversible; the symptoms usually disappear after administration of calcium. The halflife of therapeutic doses of streptomycin is generally about 5 h in adults. Tinnitus is a good indicator of acute damage to the auditory system. Streptomycin can also produce dysfunction of the optic nerve. Irritation and sensitization  Streptomycin is known to induce allergic reactions. Examples include a cattle breeder who developed allergic contact dermatitis during repeated handling of the antibiotic (Gauchía et al., 1996). In other cases reported by Pérez et al. (1996), two women with malignant melanoma received subcutaneous immunotherapy treatment and subsequently presented with local paresthesia and erythematous lesions or pruritic exanthema. The sensitization was traced to streptomycin sulfate routinely added to cell culture media of the killer cells used for the immunotherapy. Effects on reproduction  Streptomycin can cause certain congenital malformations as well as hearing loss and deficient vestibular functions in children born to women who received the drug during the first trimester of their pregnancy (Donald et al., 1991; Holdiness, 1987). Poisoning incidents  There are only a few reported acute streptomycin poisoning incidents. A recent fatal case concerned intravenous drug abuse involving the injection of a combination of streptomycin and penicillin G (Rance and Randall, 1986). Death was apparently caused by streptomycininduced neuromuscular blockade. The blood streptomycin level was 207 mg/l. Several intravenous injection sites at the inguinal areas were found. Pathology revealed diffuse inguinal and axillary lymphadenopathy that microscopically showed reactive lymphoid hyperplasia. Relatives had also noted a history of progressive hearing loss and episodes of partial paralysis. Recently Romano et al., (2002) (see also Iikura et al., 2002) have described a farmer who suffered an anaphylactic shock following an accidental contact with streptomycin. The antibiotic absorbed through skin lesions triggered within minutes angioedema of the hand, generalized urticaria and angioedema with dyspnea, dizziness, and severe hypotension. After treatment that included intramuscular epinephrine, intravenous methylprednisolone and chlorpheniramine, the symptoms resolved in 10 h. Therapy  Calcium salts are known inhibitors of the uptake and binding of aminoglycosides.

Chapter | 3  Pest Control Agents from Natural Products

3.4  Herbicides Although there has been intensive research to find practically useful weed control agents from natural sources (Duke, 1986; Duke et al., 2002), only two related compounds of microbial origin have been commercialized.

3.4.1  Bilanafos (Bialaphos) (a)  Introduction Bilanafos, originally called bialaphos (Figure 3.13), was isolated from the culture filtrates of the actinomycetes Streptomyces viridochromogenes (Bayer et al., 1972) and Streptomyces hygroscopicus (Kondo et al., 1973; Tachibana, 2003; Tachibana and Kaneko, 1986). This unique phosphorus-containing tripeptide-like antibiotic is now used as a nonselective postemergent herbicide. (b)  Identity, Physicochemical Properties, and Uses IUPAC name: 4-[hydroxy(methyl)phosphinoyl]-l-homoala­ nyl-l-alanyl-l-alanine. Chemical Abstract name: (2S)-2-amino-4-(hydroxymethyl phosphinyl)butanoyl-l-alanyl-l-alanine. CAS Registry Numbers: bilanafos: [35597-43-4]; bilanafos-sodium: [71048-99-2]. The empirical formula of bilanafos is C11H22N3O6P, and the molecular weight is 323.3. The empirical formula of bilanafos-sodium is C11H21N3O6PNa, and the molecular weight is 345.3. Physicochemical properties  The solubilities of bilanafos in water, methanol, and ethanol are 1000, 500, and 250 g/l, respectively. It is practically insoluble in acetone, chloroform, and ethyl ether. Bilanafos-sodium is soluble in water and methanol, but insoluble in ethanol, acetone, ethyl ether, chloroform, and hexane. Bilanafos is levorotatory: [ ]25 D  34 (c  10% in water). Stability  Bilanafos is degraded microbially in the soil through phosphinothricin [35597-44-5] (Figure 3.13) with

P OH

O

O

O

NH2

H N

N H

O

bilanafos

O

O

P OH

OH NH2

phosphinothricin

Figure 3.13  Structures of microbial herbicides.

OH

183

a half-life of about 5 h (Sekizawa and Takematsu, 1983), though in different soils half-lives of 10 days or longer have been reported (Jobidon, 1991). Formulations and uses  Bilanafos or its sodium salt is formulated as soluble powder and sprayable liquid. It is used at 0.2–3 kg/ha application rates in vines, orchards, rice and other crops as well as to control weeds in uncultivated lands. (c)  Biological Properties Mode of action  Bilanafos is actually a proherbicide. The tripeptide has no activity in vitro, but in microorganisms and plants it is converted to the actual bioactive agent phosphinothricin by nonspecific intracellular peptidases. Phosphinothricin, as a glutamic acid analog, irreversibly inhibits glutamine synthetase (EC 6.3.1.2) leading to a decrease in the amino acids glutamine, aspartate, serine, glycine, and alanine in the affected organism. The herbicidal effect can be reversed by glutamine (see Bayer et al., 1972). Inhibition of the enzyme also causes accumulation of the cell toxicant ammonia (Tachibana et al., 1986) and impaired photosynthesis (Wendler et al., 1992). The biochemistry and genetics of bilanafos were reviewed by Thompson and Seto (1995). (Phosphinotricin is discussed in more detail in Section 3.4.2.) Metabolism and excretion  Suzuki et al. (1987) studied the metabolism and excretion of orally administered [14C]bilanafos in mice. An almost complete degradation of bilanafos was demonstrated with over 89% of the radioactivity excreted into the feces and about 8% of the absorbed radioactivity excreted into the urine within 24 h. Among the three metabolites detected in the feces, phosphinothricin was the predominant one (50%). (d)  Toxicity to Laboratory Animals Acute and chronic toxicity  The oral acute LD50 values of bilanafos-sodium in male and female rats are 268 and 404 mg/kg, respectively. The acute oral LD50 value for chicken is greater than 5000 mg/kg. For carp and Daphnia magna the 48-h LC50 value is 1000 ppm for each. Bilanafos was not mutagenic in the Ames assay (Tomlin, 2003). Poisoning syndromes  The symptoms of acute oral bilanafos poisoning in rats are hypothermia, cyanosis, multidirectional nystagmus, apnea, and convulsions with paroxysmal waves in EEG. The cause of death is respiratory arrest and the pathological findings can show hemorrhage in the digestive tract, hematuria, hepatic atrophy, and adrenal swelling (see Matsukawa et al., 1991). (e)  Toxicity to Humans Poisoning incidents  Matsukawa et al. (1991) described a bilanafos poisoning case. A man under alcoholic intoxication

184

ingested 100 ml of a 32% liquid formulation of bilanafos (Herbiace) and started to vomit shortly. Serum phosphinothricin concentration, measured on admission to the hospital within an hour of ingestion, was 15 mg/l, body temperature was 34.8°C, blood pressure was 108/60 mmHg, the heart rate was 108/min, and the lips and nails were cyanotic and glycosuria was noted. Gastric lavage, enema, orogastric intubation, and urinary catheterization were performed. Infusion of saline, sodium bicarbonate, and furosemide corrected acidosis by 15 h. At 10 h after ingestion, nystagmus was observed, which lasted for 19 days. At 40 h, apnea and convulsions developed progressively and normal respiration returned only at 108 h after ingestion. EEG signs were atypical triphasic waves and slow waves. In another poisoning case with similar symptoms Ohtake et al. (2001) applied hemodialysis and hemoperfusion to remove bialaphos and phosphinothricin from initial concentrations of 0.33 g/ml to 0.05 g/ml and from 14 g/ml to 0.86 g/ml, respectively, within 3 days. In accordance with the glutamine synthetase inhibitory action of the herbicide, analysis of plasma and cerebrospinal fluid indicated decreased glutamine and increased glutamate concentrations. Treatment  No specific antidote exists for bilanafos poisoning. Respiratory support, hemodialysis, and treatment of acidosis are important. (See also glufosinate poisoning in the following section.)

3.4.2  Glufosinate (a)  Introduction Glufosinate is a nonselective contact and systemic herbicide. It is a racemic mixture of phosphinothricin (Figure 3.13) and its stereoisomer. In fact, phosphinothricin is the actual bioactive metabolite of the related herbicide, bilanafos (see above) (Bayer et al., 1972). Glufosinate is not a genuine natural product because it is manufactured by chemical synthesis and commercialized as the ammonium salt, glufosinate-ammonium (GLA). The biological properties of the herbicide, including mode of action and toxicology, were reviewed (Donn, 2007; Hoerlein, 1994). (b)  Identity, Physicochemical Properties, and Uses IUPAC name: glufosinate: 4-[hydroxy(methyl) phosphinoyl]-dl-homoalanine. Chemical Abstract name: phosphinothricin: (2S)-2amino-4-(hydroxymethylphosphinyl)butanoic acid; glufosinate-ammonium: ammonium ()-2-amino4-(hydroxymethylphosphinyl)butanoate. CAS Registry Numbers: phosphinothricin [35597-44-5], glufosinate: unspecified stereochemistry [51276-472]; racemic [53369-07-6]; glufosinate-ammonium [77182-82-2].

Hayes’ Handbook of Pesticide Toxicology

Empirical formula: glufosinate: C5H12NO4P; molecular weight: 181.1; glufosinate-ammonium: C5H15N2O4P; molecular weight: 198.2. GLA is a slightly pungent, crystalline solid that melts at 215°C. The solubility of the ammonium salt at 22°C in water is greater than 1370 g/l; at 20°C in acetone, 0.16 g/l; in ethanol, 0.65 g/l; in ethyl acetate, 0.14 g/l; and in hexane, 0.2 g/l. The pKa1, pKa2, and pKa3 values of glufosinate are 2, 2.9, and 9.8, respectively. The log P value of GLA is less than 0.1. Formulations and uses  The commercial herbicide is formulated as a water-soluble concentrate containing GLA, the anionic surfactant sodium polyoxyethylene alkylether sulfate (SPAS), and propylene glycol ether. It is used to control a wide range of annual and perennial broad-leaved weeds and grasses in orchards, rubber and oil palm plantations, ornamental trees, and noncroplands. It is also used as a desiccant in potatoes and sunflowers. Application rates are 0.4–1.5 kg/ ha. Glufosinate-resistant transgenic crop varieties, including canola, cotton, maize, rice, soybean, and sugarbeet that tolerate the nonselective but rapidly degradable herbicide were recently developed (Donn, 2007; Tan et al., 2006). (c)  Biological Properties Mode of action  Glufosinate has the same herbicidal mode of action as bilanafos or phosphinothricin (see above). Studies with glufosinate indicated irreversible inhibition of wheat glutamine synthetase, a key plant enzyme of the assimilation of inorganic nitrogen (Manderscheid and Wild, 1986). Once phosphorylated at the target site, the herbicidally active l-glufosinate (or S-isomer, i.e., phosphinothricin) is bound to the enzyme leading to a rapid increase of ammonium concentration, a deficiency in glutamine, and inhibition of photorespiration and photosynthetic processes that are ultimately responsible for phytotoxicity. Recently, the x-ray crystal structure of the maize enzyme complexing phosphinothricin has been solved (Unno et al., 2006). Glufosinate has been shown to possess acaricidal (Ahn et al., 1997), insecticidal (Kutlesa and Caveney, 2001), antibacterial (Pline et al., 2001), and fungicidal (Albrecht and Kortekamp, 2008) activities. Glutamine synthetase is also present in vertebrate CNS and the mouse enzyme has been shown to be inhibited by phosphinothricin at micromolar concentrations in vitro (Lapouble et al., 2002). Glufosinate and phosphinothricin are phosphinic acid analogs of glutamate, which is an important excitatory neurotransmitter in animals (reviewed by Bak et al., 2006). Metabolism, excretion, and degradation  The metabolism and the environmental fate of glufosinate were reviewed by Hoerlein (1994). In all test species, including rats, dogs, goats, and hens, 80–90% of an oral dose

Chapter | 3  Pest Control Agents from Natural Products

of glufosinate was excreted unchanged in the feces over 48 h, whereas 10–15% was eliminated in the urine. The typical metabolites in rat feces were N-acetylglufosinate and -hydroxy-4-(methylphosphinyl)butyric acid (MHB) amounting to 7–9 and 3–4% of the administered dose, respectively. The rat urinary metabolites identified were 4-(methylphosphinyl)butyric acid (MPB) and 3-(methyl­ phosphinyl)propionic acid (MPP) (2% each). Both isomers of glufosinate are readily degraded microbially in soil by initial transamination to the corresponding 2-oxoacid, which then undergoes further transformations eventually leading to 2-(methylphosphinoyl)acetic acid and CO2 (Bartsch and Tebbe, 1989; Smith, 1988; see also Faber et al., 1998). At 20°C, the half-lives of the herbicide in various soils were 3–7 days, as reported by Gallina and Stephenson (1992). Other studies (see U.S. EPA, 2008d) found GLA to be more stable to abiotic degradation processes; for example under anaerobic soil metabolism conditions the half-life of the herbicide was 56 days. Recent microcosm studies by Pampulha et al. (2007) have indicated a long-lasting, and in some cases inconsistent, influence of GLA or its metabolite(s) on the population and enzymatic activity of soil fungi and bacteria. In genetically modified glufosinate tolerant crops, the herbicide is converted into its inactive N-acetate derivative by the phosphinothricin-N-acetyltransferase enzyme, which is originally encoded by the pat gene from Streptomyces species. In transgenic plants expressing pat, the detoxification of phosphinothricin is thus exclusively N-acetylation, while in nonresistant plants the metabolism proceeds through transamination to the corresponding -oxocarboxylic acid, which is then further transformed either by oxidative decarboxylation to propionic acid derivative MPP or by reduction to the -hydroxybutyric acid derivative MHB both retaining the terminal phosphinyl moiety (DrögerLaser et al., 1994). A recent report by Hori et al. (2003) indicated that the stereoisomers of glufosinate could be metabolized or transported differently in humans. Analysis of the cerebrospinal liquid and blood plasma samples of a suicidal poisoning case revealed that 27 h after ingestion of about 18.5 g dlglufosinate, the d-isomer predominated in these body fluids: in the CNS the isomer concentrations were 0.48 g/ml for the d-isomer and 0.12 g/ml for the l-isomer, while in the blood the concentrations were 1.44 g/ml for the d-isomer and 0.35 g/ml for the l-isomer. (d) Toxicity to Laboratory Animals Acute and chronic toxicity  Results of various toxicity studies with GLA are summarized in Table 3.13. Although GLA presents only negligible hazard, the anionic wetting agent in the commercial formulation (Basta) increases toxicity on dermal application. No adverse effects were seen in earthworms at 1000 l/ha application of 20% Basta (Hoerlein, 1994). GLA is not hazardous to bees

185

(LD50  100 g/bee) (Tomlin, 2003). Again, formulated GLA is more toxic to aquatic organisms than the active ingredient alone (see examples in Table 3.13). The inconsistency between poisoning incidents with the formulated herbicide and the reported low toxicity of the pure active ingredient prompted Koyama et al. (1997) to examine GLA and the SPAS surfactant for their cardiovascular effects in rats in vitro and in vivo. Whereas GLA had no effect on isolated atria and aortas, both the herbicide formulation and SPAS produced negative chronotropic responses in isolated atria and exerted significant vasodilative activity in phenylephrine-pretreated aortic ring segments. Intravenous administration of either the herbicide formulation or SPAS at 0.3–30 mg/kg reduced blood pressure in a dose-dependent manner. Additional symptoms noted were a slight increase in heart rate for the low doses and a marked decrease at the 30 mg/kg dose. By contrast, GLA failed to produce any of these effects strongly suggesting that the observed hypotension is caused by the surfactant of the commercial formulation. In rodents, acute intraperitoneal doses (30–100 mg/kg) of GLA provoked stereotyped tonic-clonic seizures attributed to glutamatergic effects, specifically the activation of NMDA receptors (Lapouble et al., 2002; Matsumura et al., 2001). Chronic exposure to low doses of the herbicide did not affect locomotor activity but induced memory impairment, modified hippocampal texture, and increased hippocampal glutamine synthetase activity (Calas et al., 2008). Examining the acute toxicity of a GLA water-soluble formulation (24.5% active ingredient) to amphibian species, Dinehart et al. (2009) have found that a 48-h exposure of juvenile Great Plains toads, Bufo cognatus, and New Mexico spadefoots, Spea multiplicata, to a surface covered with paper impregnated with the herbicide at 0.21 ml/m2 resulted in 75 and 74% survival, respectively. Exposure to soil treated at the same rate, the 48-h survival was 100% for each species. The N-acetylglufosinate metabolite is of low acute toxicity in rats (oral LD50  290 mg/kg, intraperitoneal LD50  1200 mg/kg), not genotoxic and not a rodent carcinogen (World Health Organization, 2000). Toxicity symptoms and pharmacology  Hack et al. (1994) reported the results of a detailed mammalian pharmacological and neurotoxicological study. Intracerebroventricular administration of 10 g GLA to male rats elicited slight spasms of the forelimbs and opisthotonos (arched-back body), which was responsive to diazepam (10 mg/kg). A 20-g dose produced general convulsions after a latency period of approximately 3 h. The symptoms could be alleviated by an intraperitoneal injection of diazepam but recurred within 24 h. An analysis of catecholamine levels in various parts of the brain revealed a significant increase in dihydroxyphenylacetic acid level in the striatum and a decrease in the norepinephrine level of the frontal cortex for

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Table 3.13  Acute Toxicity of Glufosinate-ammonium and its Formulated Products Species, sex

Route

LD50 (mg/kg)

Rat, male/female

oral

1660/1510

Other data

References Ebert et al. (1990)

a

Rat, male/female

oral

2170/1910

Rat, male/female

sc

73/61

Ebert et al. (1990)

Rat, male/female

ip

96/83

Ebert et al. (1990)

Rat, male/female

dermal

4000/4000

Rat, male/female Rat, male/female

dermal inhalation, 4-h

1400/1380

a

1400/1380

a a

Rat, male/female

inhalation, 4-h

1400/1380

Mouse, male/female

oral

436/464

Ebert et al. (1990)

Ebert et al. (1990) Ebert et al. (1990) b

World Health Organization (1992)

LC50  0.621 mg/l

c

LC50  1.26/2.60 mg/l

World Health Organization (1992) Ebert et al. (1990)

a

Mouse, male/female

oral

1420/1570

Mouse, male/female

sc

88/104

Ebert et al. (1990)

Mouse, male/female

ip

103/82

Ebert et al. (1990)

a

Rabbit

oral

Mallard duck

8-day, dietary

LC50  5000 ppm

U.S. Department of Agriculture and Office of Pesticide Programs (2009)

Trout

96-h

LC50  710 ppm

Hoerlein (1994)

Trout

96-h

LC50  15 ppma

Hoerlein (1994)

Bluegill

96-h

LC50  320 ppm

Hoerlein (1994)

Bluegill

96-h

LC50  56–75 ppma

Hoerlein (1994)

Daphnia magna

48-h

LC50  560 ppm

Hoerlein (1994)

Daphnia magna

48-h

LC50  15 ppma

Hoerlein (1994)

Daphnia magna

48-h, static

LC50  668 ppm

USDA OPP (2009)

Daphnia magna

21-day, static renewal

LC50  56 ppm

USDA OPP (2009)

Eastern oystern

48-h

LC50  8.0 ppm

Eastern oystern

1550

Ebert et al. (1990)

48-h

Green algae Zooplankton community

Ebert et al. (1990)

USDA OPP (2009)

a

USDA OPP (2009)

LC50  2.7 ppm

Hoerlein (1994)

LC50  37 ppm 6-day

d

EC50  0.24 ppm

Faber et al. (1998)

a

A 200 g/l glufosinate-ammonium liquid formulation containing anionic surfactant, propylene glycol ether, defoamer, dye, and water. Aerosol. c Dust. d A 137 g/l glufosinate-ammonium liquid formulation. b

animals receiving the 20-g dose. No effects were seen for the 10-g dose treatment. Brain glutamine synthetase activity, however, showed slight to moderate dose-dependent inhibition not only after intracerebroventricular but after intravenous (100 mg/kg) application as well. Oral application of a single dose of 1600 mg/kg of glufosinate to female rats caused poisoning symptoms, starting with diarrhea 6 h after treatment. Convulsions,

restlessness, and piloerection were also observed. Intoxication reached a maximum on days 2 and 3 after treatment, with tonoclonic convulsions, squatting position, lagophthalmos, drowsiness, reduced respiration, and bloodencrusted eyelids and snouts. Some of the animals succumbed. At lower doses, the symptoms were less severe. The signs of intoxication receded 3.5 days after treatment. Furthermore, a decrease in glutamate synthetase activity in

Chapter | 3  Pest Control Agents from Natural Products

the liver and the kidneys was noted. In the kidneys, enzyme inhibition was detectable 4 h after dosing, was maximum on day 1, and disappeared completely within 7.5 days; in the liver, it was more lasting. The brain enzyme was less sensitive. The glutamate level slightly increased in the liver and decreased in the brain. GLA had no effect either on Ca2 channels from rat frontal cortex or on GABA, benzodiazepine, norepinephrine, dopamine, and serotonin rat or bovine receptors in vitro. A circadian-stage-dependent mouse toxicity study by Yoshiyama et al. (1995) found that mortality was highest when the animals received the GLA formulation (Basta) (oral 1500 and 3000 mg/kg doses) at the beginning of the light-on phase; mortality was lowest for administration at the beginning of the dark phase. Mutagenicity, oncogenicity, and reproductive toxicity  GLA was nonmutagenic in bacteria, yeast, and in vitro and in vivo mammalian genotoxicity assays. There was no evidence for oncogenic potential in mice in a 2-year study with a maximum of 160 ppm (males) or 320 ppm (females) dietary concentration of the herbicide. In a two-generation reproductive toxicity study with rats, the NOEL for fertility and reproductive performance was 120 ppm dietary GLA, equivalent to an average daily dose of 12 mg/kg for the dams during pregnancy and lactation. In embryotoxicity studies, maternal toxicity was noted for rats at the highest doses of 50 and 250 mg/kg, whereas maternal toxicity for rabbits occurred at the highest dose of 20 mg/kg. The herbicide was not teratogenic (Ebert et al., 1990). GLA caused growth retardation, various morphological abnormalities, and lethality in developing mouse embryos in vitro at 105 M concentrations and in cultured embryonal cells of the midbrain and the limb bud at 106 M concentrations (Watanabe, 1997; Watanabe and Iwase, 1996). GLA specifically affected the neuroepithelium of the brain vesicle and neural tube, leading to apoptosis by an unknown mechanism; nevertheless, the observed excitotoxic cell death was similar to that caused by glutamate observed in other studies. Assessing the reproductive effects of glufosinate observed in the laboratory, Schulte-Hermann et al. (2006) have concluded that the effects found in rats occur only at levels which would not be experienced under either occupational or nonoccupational exposures under normal handling or use, thus data demonstrating impaired preimplantation or implantation in laboratory animals have no relevance to humans. (e)  Toxicity to Humans Poisoning incidents and treatments  Suicidal ingestions of GLA formulations caused respiratory failures and delayed nervous system disorders with a mortality rate of 19% in Japan (Koyama, 1999). In general, the pathophysiology of human GLA poisoning can be classified into two

187

categories: initial gastrointestinal symptoms due to the irritant effect of surfactant in the herbicide and to the neurological and circulatory failures occurring after a latency period of about 1 day and are related to the direct or indirect effects of GLA. Koyama et al. (1994) described a poisoning case of a woman ingesting 500 ml of Basta containing 18.5% GLA plus 35% anionic surfactant. Immediately after ingestion, she developed nausea and vomiting, which subsided within a few hours. Treatment consisted of gastric lavage, administration of charcoal and a cathartic, and forced diuresis with furosemide. The patient, however, gradually lost consciousness and was in deep coma with general cyanosis 9 h after ingestion. Intubation followed by artificial ventilation was initiated. Hemoperfusion for 4 h did not improve clinical signs. Consciousness and spontaneous respiration were slowly regained and extubation was possible on day 8. The patient had generalized edema from days 1 to 5 and elevated body temperature up to 40°C from days 1 to 8. Endoscopy showed erosion of the gastric mucous membranes. No convulsions developed in this case. Toxicity was attributed to the anionic surfactant in the herbicide formulation. Based on this and ten other poisoning cases, the human acute oral toxic dose that causes delayed consciousness disturbance was estimated to be 1.6–1.8 ml/kg Basta (corresponding to 296–333 mg/kg GLA) (Koyama et al., 1995). Tanaka et al. (1998) reported two suicidal ingestion cases, which were complicated by general convulsions that developed 8.5 and 33 h after ingestion. Watanabe and Sano (1998) described in detail the suicidal poisoning case of a man who ingested 180 ml herbicide formulation (corresponding to 33.3 g GLA). Symptoms, developing shortly after ingestion, included vomiting, diarrhea, and impaired consciousness. Upon hospitalization, metabolic acidosis, a body temperature of 35.4°C, a pulse rate of 110 beats/min, and no detectable diastolic blood pressure were determined. Emergency treatment consisted of intubation and gastric lavage, followed by the administration of charcoal, diuretics, and a purgative. White blood cell, glucose, and urea nitrogen levels were elevated during the first 5 days, whereas lactate dehydrogenase activity peaked on day 4 of hospitalization. The cholinesterase level was reduced during the first 5 days. The urine glufosinate level was higher than 40 g/ml on day 2 but undetectable on day 3 after direct hemoperfusion. The convulsion occurring on day 2 was responsive to thiopental sodium and diazepam. Magnetic resonance imaging demonstrated slight ischemia-induced changes in the white matter of the lateral brain regions. During recovery, retrograde and anterograde amnesia appeared. Park et al. (2006) have recently described an accidental poisoning case. In addition to recurrent nausea and vomiting, the victim was mentally confused and showed increased irritability. Treatment involved gastric lavage followed by charcoal. He was released with no medical

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188

problems a week after admission. However, 10 days after discharge, the patient consulted the hospital because of anterograde amnesia. Neuropsychological tests indicated memory dysfunction and learning difficulties, apparently due to delayed neurotoxic effects. Magnetic resonance imaging of the brain revealed bilateral hippocampal lesions considered to be responsible for the experienced anterograde amnesia; complete remission was seen 6 months later. Apparently, GLA intoxication caused excessive stimulation of NMDA receptors leading to reversible excitotoxic cell damage in the hippocampus of the patient. Additional suicidal poisoning cases with progressive recovery have been described (Lawson and EstradeChapellaz, 1999; Lluís et al., 2008; Takahashi et al., 2000). Treatment  Treatment is generally symptomatic as illustrated above. Activated charcoal is expected to adsorb the surfactant but not GLA in the herbicide formulation. Diuresis is believed to be essential, but the clinical usefulness of hemopurification remains to be demonstrated (see below). It is essential to monitor vital signs closely during hospitalization for several days after poisoning regardless of the amount ingested, to provide respiratory support, and to expect delayed symptoms, even when the primary treatment is complete (Koyama, 1999; see also Hori et al., 2003). Based on a study with bovine blood contaminated with 268–280 ppm herbicide formulation, Tanaka et al. (1995) preferred hemodialysis to direct hemoperfusion because the former removed over 99% of GLA, whereas the latter extracted merely 3% of GLA within 2 h; the removal of the surfactant, however, was not examined. The efficacy of hemodialysis and hemoperfusion in removing GLA was also demonstrated in a human poisoning case with initial serum concentration of 1.79 g/l GLA (Shinohara et al., 1997).

Some of the rodenticides surveyed in this chapter appear to have become obsolete; nevertheless, their unique toxicology and the fact that, from time to time, they still resurface in some countries justify a brief discussion. General reviews on the history of rodent control (Chitty, 1954) and on the chemistry of important natural and synthetic rodenticides (Elliott, 1995) are available.

3.5.1  Strychnine (a)  Introduction The extremely poisonous alkaloid strychnine (Figure 3.14) was isolated in pure form by Pelletier and Caventou in 1818 from St. Ignatius beans, Strychnos ignatii (Loganiaceae), a woody vine native to the Philippines. It is now obtained from the ripe and dried seeds of S. nux vomica, a related plant growing in India, Sri Lanka, and Southeast Asia. The seeds contain 1.0–1.5% strychnine and about the same amount of its 2,3-dimethoxy derivative, brucine. Philippe et al. (2004) have recently reviewed the ethnobotany, pharmacotoxicology, and chemistry of various Strychnos alkaloids. (b)  Identity, Physicochemical Properties, and Uses IUPAC and Chemical Abstract name: strychnidin-10-one. CAS Registry Number: [57-24-9]. Empirical formula: C21H22N2O2; molecular weight: 334.4. 21

N

22

H

H O

2

H

N

3

O strychnine

3.5  Rodenticides Rodenticides are used to kill mammals that compete for our food, are vectors of fatal diseases such as rabies and the plague, and damage buildings, dams, or underground cables. Rodent control agents occupy a unique place among pest control agents due to their high vertebrate toxicity and also because these poisons have been among the most frequently misused pesticides (Barnett and Fletcher, 1998; Parsons et al., 1996). While inadequate safety and effectivity of such lethal chemicals have severely restricted their use, ethical considerations have emerged recently dictating the avoidance of unnecessary suffering of the target vertebrate pests (see Meerburg et al., 2008). The management of resistance to synthetic anticoagulant rodenticides requires alternative substances. In the case of an outbreak of disease, single-dose toxicants of natural origin, such as strychnine and red squill, provide economic and rapid reduction in the rodent population.

O O

OH

R

3 O

H

OH

6 O O

scilliroside R = β-D-Glc Figure 3.14  Structures of rodenticides.

Chapter | 3  Pest Control Agents from Natural Products

Physicochemical properties  Strychnine is an odorless base (pKa  8.26) forming, in pure form, colorless or white crystals that melt at 275–285°C (with decomposition). The solubility of the free alkaloid and its sulfate in water is 143 mg/l at ambient temperature and 30 g/l at 15°C, respectively (Tomlin, 2003). The solubility of the free base in ethanol and chloroform is 6.7 and 200 g/l, respectively. The log P of the free base is 4.0 at pH  7. Strychnine is levorotatory: [ ]25 D  139 (c  0.4 in chloroform). Strychnine is very bitter with a taste threshold of 1.4 ppm in solution (Budavari, 1996). A dilute solution of strychnine in 80% sulfuric acid gives a reddish-violet to bluish-purple color on the addition of a trace amount of potassium dichromate solution (Otto reaction). Under abiotic conditions, strychnine is a relatively stable compound. It is photostable and does not hydrolyze at pH 5–9. The alkaloid is immobile in the soil where degradation is entirely microbial (Rogers et al., 1998b; U.S. EPA, 1996). History of use and formulations  The use of the seeds of S. nux vomica as a rat poison was introduced in Germany in the late 17th century. Since then, strychnine, in one form or another, has been used worldwide to kill vertebrate pests, including moles, skunks, gophers, mice, rabbits, coyotes, and various predators, as well as sparrows, pigeons, and other unwanted birds. The wild bear population could also be effectively controlled by strychnine nitrate baits (Inukai, 1969). Due to its bitterness, however, the alkaloid is not suitable as a rat poison (bait shyness). Typical formulations in baits are pellets, grain, or eggs containing 0.25–1% of the alkaloid. Strychnine is a single-dose, acute toxicant. In the United States, strychnine-containing products are classified as restricted use for belowground applications such as for gopher control (U.S. EPA, 1996). In the European Union, use of strychnine in agriculture and rodent control has been banned as of September 2006. Strychnine has been listed in pharmacopoeias of many countries as a tonic and stimulant in veterinary and human medicine, though its use as a doping agent is forbidden. Homeopathic S. nux vomica preparations are also widely available as over-the-counter products. (c)  Biological Properties Mode of action  Strychnine is a strong convulsant. The alkaloid excites the CNS by antagonizing the effects of the inhibitory neurotransmitter, glycine (reviewed by Lynch, 2004). The strychnine-sensitive glycine receptor (GlyR) is a ligand-gated chloride ion channel consisting of five subunits embedded in the postsynaptic membrane. Binding of glycine results in opening of the channel and increased Cl conductance across membranes resulting in hyperpolarization and inhibition of the postsynaptic neuron. Glycine and strychnine bind competitively to an overlapping but

189

not identical binding site at the GlyR protein, and this provides the structural basis of the antagonist behavior of the alkaloid. The three-dimensional structure of GlyR is not known but it shows structural and functional similarities to nAChRs. (Note, that glycine is also a co-agonist of excitatory NMDA-type glutamate receptors on cation-selective ion channels but these sites are not affected by strychnine.) The distribution of inhibitory GlyRs in the human brain was mapped by using [3H]strychnine (Probst et al., 1986). The receptors are most abundant in the spinal cord and brain stem where they are mainly involved in processing motor and sensory information. Strychnine-binding GlyRs were also found in the cortex, the auditory system, and the retina. Thus, when inhibition is blocked, ongoing neuronal excitability is increased and sensory stimuli produce exaggerated reflex effects. Higher brain centers such as the substantia nigra, neostriatum, and hippocampus are also relatively insensitive to strychnine, explaining why poisoning symptoms are largely spinal in origin. Human startle disease (hyperekplexia), a rare hereditary neurological disorder characterized by an exaggerated reaction (for example, anxiety and muscular rigidity) to unexpected stimuli, has been associated with mutation of a GlyR subunit. Strychnine can also depress nicotinic–cholinergic responses through interaction with nicotinic receptors and at high concentrations in vitro binds to other receptors as well (see, for example, Jensen et al., 2006). In rodent spinal cord, there are two major GlyR isoforms. The receptor variant of newborn rodents is a homopentamer made up of polypeptides with a molecular weight of 49 kDa (Becker et al., 1988). The adult GlyR isoform, however, is a complex glycoprotein consisting of three polypeptides with molecular weights of 48, 58, and 93 kDa (Pfeiffer et al., 1982). The neonatal receptor is predominantly expressed around birth and has low strychninebinding affinity; within 2 weeks after birth, it is replaced by the adult receptor form, which is strychnine sensitive (Becker et al., 1988; Brüning et al., 1990). Absorption, metabolism, and excretion  Strychnine is rapidly absorbed from the gastrointestinal tract and nasal mucosa but not through the skin. Symptoms begin about 15–60 min following ingestion, delayed presentations have rarely been reported. Distribution of the drug in tissues is also rapid as is its metabolism to several nontoxic polar products by hepatic enzymes (Adamson and Fouts, 1959; Oguri et al., 1989). Only 5–20% of the intact alkaloid is excreted in urine. The metabolism of the alkaloid was inhibited by the CYP blocker SKF-525 in rodents (Adamson and Fouts, 1959; Kato et al., 1962) but was induced by phenobarbital (Kato et al., 1962). The different oral toxicities of strychnine to guinea pigs and to rats were attributed to different metabolic rates in these rodents (Kato et al., 1963; see Table 3.14). The observation that

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190

Table 3.14  Acute Toxicity of Strychnine to Laboratory Animals Species

Route

LD50 (mg/kg)

Note

References

Rat, male

iv

0.57

Kato et al. (1962)

Rat, female

iv

0.57

Kato et al. (1962)

Rat, male

ip

2.82

Kato et al. (1962)

Rat, female

ip

1.62

Kato et al. (1962)

Rat, male

sc

4.01

Kato et al. (1962)

Rat, female

sc

1.81

Kato et al. (1962)

Rat, male

ip

3.1

Blum and Zacks (1958)

Rat

ip

1.5

at 26°C

Keplinger et al. (1959)

Rat

ip

0.25

at 8°C and 36°C

Keplinger et al. (1959)

Rat

iv

0.96

Setnikar and Magistretti (1967)

a

Rat

sc

2.73

Kamel and Afifi (1969)

Rat, male

oral

6.4

U.S. EPA (1996)

Rat, female

oral

2.2

U.S. EPA (1996)

Mouse

oral

8.0

Schafer and Bowles (1985)

Mouse, male

sc

1.45

Kretzschmar et al. (1970)

ip

b

Lakatos et al. (1964)

b

Lakatos et al. (1964)

Mouse, male Mouse, female Guinea pig, female Guinea pig, female Guinea pig, female Rabbit Mongrel dog

ip ip

1.9

1.6

10.9

b

Kato et al. (1963)

b

Kato et al. (1963)

iv

0.39

sc

b

dermal sc

4.8

Kato et al. (1963)

2000

c

a

0.46

U.S. EPA (1996) Kamel and Afifi (1969)

a

Strychnine hydrochloride. Strychnine sulfate. c No signs of toxicity observed. b

female rats were more susceptible to strychnine than males (Poe et al., 1936) was explained by more efficient hepatic metabolism in the latter sex (Kato et al., 1962). The oxidative nature of the metabolism was demonstrated in rats where strychnine 21,22-epoxide and strychnine N-oxide were identified as the respective major and minor urinary metabolites (Oguri et al., 1989). In nonfatal human poisoning cases, strychnine elimination followed first-order kinetics with half-lives of 10–16 h (Edmunds et al., 1986; Palatnick et al., 1997; Wood et al., 2002). Studies indicate that the use of underground strychnine baits to suppress pocket gophers poses minimal risk to nontarget species, including insects that feed on carcasses of the poisoned rodents (Arjo et al., 2006; El Hani et al., 2002; Ramey et al., 2002). In connection with a major

mouse plague control program in South Australia during 1993, degradation of strychnine by various soil microbes has been studied (Rogers et al., 1998a, 1998b). (d) Toxicity to Animals Acute toxicity  The toxicity of strychnine and its salts was thoroughly studied, and representative acute toxicity data are listed in Tables 3.14 and 3.15 for experimental and wildlife animals, respectively. For additional data, see Ray (1991). The acute toxicity of strychnine to rats was shown to be influenced by environmental temperature (Keplinger et al., 1959) and altitude (Moore and Ward, 1935). The poisoning syndrome in animals is essentially the same as that observed in humans and is described in the following section (see also Meiser and Hagedorn, 2002).

Chapter | 3  Pest Control Agents from Natural Products

191

Table 3.15  Toxicity of Strychnine to Wildlife Species Species

LC50 (ppm)

Other data

References

LD50  5.0a

Schafer and Bowles (1985)

Acute toxicity Starling (Sturnus vulgaris) (oral) European ferret (Mustella putorius) (dietary, 5-day)

198

Striped skunk (Mephitis mephitis) (dietary) Red fox (Vulpes fulva) (dietary, 5-day)

U.S. EPA (1996) LD100  31 mg/ egg/skunk

70

U.S. EPA (1996) Lethal dose  0.5 mg/kgb

Bear (Ursus arctos yesoensis) (acute oral) Subacute dietary toxicity Northern bobwhite quail (Colinus virginianus)

3536

U.S. EPA (1996)

Northern bobwhite quail (28-day)

4974

Sterner et al. (1998)

Mallard duck (Anas platyrhynchos)

212

U.S. EPA (1996)

Mallard duck

680

Sterner et al. (1998)

Black-billed magpie (Pica pica)

99

U.S. EPA (1996)

American kestrel (Falco sparverius)

234

U.S. EPA (1996)

a

Strychnine sulfate. Strychnine nitrate.

b

A recent report by Stoltenow et al. (2002) has described poisoning and treatment of six horses accidentally fed strychnine-laced barley containing at least 100 ppm of the alkaloid. Initial differential diagnosis included tetanic seizures, muscle rigidity, facial paralysis, nystagmus, and toxicosis. Two of the horses died before or during treatment that included xylazine muscle relaxant, diazepam with or without pentobarbital repeatedly administered along with initial nasogastric administration of activated charcoal slurry in water and mineral oil. Complete recovery for one of the horses required 4 weeks. Although strychnine is generally less toxic to avian species than to mammals, the way it is commonly applied poses a danger to nontarget birds (Martínez-Lopez et al., 2006; Warnock and Schwarzbach, 1995; Wobeser and Blakley, 1987). Field studies, however, found negligible tertiary risks from the use of underground baits (Arjo et al., 2006). The increasingly restricted access to the rodenticide resulted in a continuous decline of domestic and wild animal poisoning cases during the past decade (see, for example, Berny, 2007). Pathology  Autopsy findings are nonspecific and reflect only the presence of violent convulsions and anoxia.

Hemorrhages were sporadically observed in the brain of poisoned rats (Pensa and Ceccarelli, 1968), and in the myocardium and intestines of poisoned aquatic birds (Sterner et al., 1998; Wobeser and Blakley, 1987). Treatment of poisoning in animals  Because strychnine-induced death is mainly due to respiratory failure, artificial respiration will protect animals from an otherwise fatal dose of strychnine. Muscle relaxants and sedatives including diazepam, mephenesin, and barbiturates are traditionally useful drugs in treating strychnine poisoning (summarized by Ray, 1991). Kavapyrone constituents of the roots of the Polynesian kava plant, Piper methysticum, were also shown to antagonize the convulsant and lethal action of strychnine in mice (Kretzschmar et al., 1970). Upon intraperitoneal pretreatment at 300 mg/kg, methysticin (LD50  530 mg/kg) raised the subcutaneous LD50 of strychnine from 1.45 to 7.3 mg/kg, being thus more active than mephenesin and less active but safer than phenobarbital. Kavapyrones were shown to exhibit neuroprotective activity against experimentally induced ischemia in rats and mice (Backhauß and Krieglstein, 1992). However, the safety of the chronic use of kava preparations has been questioned (Teschke et al., 2008).

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Strychnine-induced tonic extensor seizures in mice were selectively and effectively blocked by a novel anticonvulsant aryltriazole derivative (MDL 27,531), administered either orally or intraperitoneally (Kehne et al., 1992). Interestingly, the compound did not affect [3H]strychnine binding in mouse brain stem and upper spinal cord membranes. Picolinic acid and its methyl ester showed anticonvulsant activity against strychnine-induced seizures in mice at 200 mg/kg intraperitoneally. These compounds also had a muscle-relaxant effect on rat decerebrate rigidity at doses of 100 mg/kg iv and depressed spinal reflexes in cats at cumulative doses of 25–200 mg/kg iv (Tonohiro et al., 1997). Milacemide (2-n-pentylaminoacetamide), an orally active anticonvulsant glycine prodrug that is able to cross the blood–brain barrier, selectively inhibited strychnineinduced allodynia in rats at doses of 100–600 mg/kg iv (Khandwala and Loomis, 1998). (e) Toxicity to Humans Acute toxicity  The typical oral lethal dose of strychnine is between 50 and 100 mg for adults and 15 and 30 mg for children; however, much higher doses have reportedly been tolerated (reviewed by Makarovsky et al., 2008; Perper, 1985; Ray, 1991). In adults, symptoms can develop from doses as low as 30 mg. Following dermal exposure, however, the prodromal period could be as long as 12 h (Greene and Meatherall, 2001). Poisoning syndrome and laboratory findings  The clinical syndrome of strychnine poisoning is very characteristic (Smith, 1990; Swissman and Jacoby, 1964). The initial symptoms, occurring as early as 15 min after oral ingestion, are mydriasis, stiffness and twitching of face and neck muscles, and movements may be abrupt. Reflex excitability is heightened and sudden tactile, visual, or acoustic stimuli induce violent motor responses. Within 30 min after ingestion, full tetanic convulsion and opisthotonos develop. The jaws are fixed (risus sardonicus) and froth gathers at the mouth. Seizures usually occur at 10- to 15-min intervals, last 30 s to 2 min, and are accompanied by loud moans of pain. During seizures, the patient remains conscious, which can arouse panic. Hyperthermia may occur and a body temperature as high as 43°C has been reported (Boyd et al., 1983). The contractions of the diaphragm and thoracic and abdominal muscles halt respiration and cause cyanosis and marked anoxia. Death is due to brain damage secondary to apnea from uncontrolled seizures (Dittrich et al., 1984) or cardiac arrest (O’Callaghan et al., 1982). For patients recovering from strychnine poisoning, lactic acidosis (arterial pH  6.55) and rhabdomyolysis with an increased creatine phosphokinase level are typical (Boyd et al., 1983). In a serious, but nonfatal case, the classical clinical poisoning syndrome was complicated by acute chemical pancreatitis (Hernandez et al., 1998b).

Hayes’ Handbook of Pesticide Toxicology

Metabolism and excretion  Within a few minutes of ingestion, strychnine appears unchanged in the urine, but the major route for removal and detoxification is oxidative hepatic metabolism (Perper, 1985; see also Boyd et al., 1983). The urine and gastric aspirate are the most useful specimens for confirming the diagnosis. Strychnine serum level as high as 2.17 g/ml at 6 h after poisoning has been reported (Hernandez et al., 1998b). Poisoning incidents  Because of the limited use of strychnine today, the number of fatal accidents caused by this nonselective poison has decreased and most of the reported cases are suicidal ingestions. There were 1000 reported cases and at least four deaths due to strychnine rodenticides in the United States between 1985 and 1990 (Klein-Schwartz and Smith, 1997). Of the 73 strychnine rodenticide poisoning cases recorded in France between 1973 and 1994, 12 were fatal (François et al., 1996). Several strychnine poisoning incidents are related to adulteration and illicit manufacture of drugs of abuse, for example, cocaine (Boyd et al., 1983; O’Callaghan et al., 1982) or heroin (Decker et al., 1982). A typical suicidal strychnine poisoning was described by Perper (1985). The fatal case involved a 79-kg man who apparently ingested about 61 g of a commercial rodenticide pellet containing 0.34% strychnine sulfate. Pathology revealed an early onset of postmortem rigidity and microscopic hemorrhages with minimal degenerative neuronal changes in the spinal cord. The mucosa of the stomach was unremarkable. The lungs were congested and edematous. Strychnine tissue concentration was highest in the bile (9.2 g/ml) and the liver (6.2 g/ml), while blood and urine contained 3.3 and 1.4 g/ml, respectively. Edmunds et al. (1986) described in detail the clinical and biochemical findings of a nonfatal suicidal attempt and summarized the clinical manifestations of 25 cases of strychnine poisoning. Among the more recent case reports (Duverneuil et al., 2004; Lindsey et al., 2004; Marques et al., 2000; Pajoumand et al., 2004; Starretz-Hacham et al., 2003; Wood et al., 2002), a poisoning with the Asian herbal remedy “maquanzi,” prepared from seeds of Strychnos species, is noteworthy (Chan, 2002). Pathology  In general, autopsy findings are nonspecific and reflect only the presence of violent convulsions and anoxia. Treatment  Emergency treatment of strychnine poisoning includes respiratory support, prevention of convulsions, and evaluation of the acid–base status. Prompt control of hyperthermia by external cooling and of convulsions with intravenous short-acting drugs such as barbiturates (e.g. pentobarbital and phenobarbital) or benzodiazepines is recommended (Boyd et al., 1983; Perper, 1985; Smith, 1990). Benzodiazepines (Herishanu and Landau, 1972) could be

Chapter | 3  Pest Control Agents from Natural Products

problematic because the doses required might also impair locomotor activity; nevertheless, the rapid-acting diazepam is the most useful anticonvulsant for this purpose because it antagonizes the convulsions without potentiating postictal depression. For severely intoxicated patients, neuromuscular blockade with pancuronium or succinylcholine (Edmunds et al., 1986) may also be needed. Mephenesin was also successful (Swissman and Jacoby, 1964). Emesis is definitely contraindicated and gastric lavage should only be done following rather than preceding the control of convulsions. For gastrointestinal decontamination, a dilute solution of potassium permanganate or, especially, activated charcoal and cathartics may be considered. Following treatment of convulsions, laboratory analysis to detect acidosis, rhabdomyolysis, myoglobinuria, and hypoxia should be performed (Edmunds et al., 1986; Hernandez et al., 1998b; Smith, 1990).

3.5.2  Red Squill and Scilliroside (a) Introduction Red squill or the sea onion, traditionally known as Urginea (Drimia or Scilla; currently Charybdis) maritima, is a large onion-like plant that grows wild in the coastal areas around the Mediterranean and is cultivated in the United States and elsewhere (Gentry et al., 1987). (The plant has been placed either in the Liliaceae or Hyacinthaceae families, depending on classification.) Its major bioactive principle, found in all parts of the plant but concentrated in the bulbs, is a bitter and emetic bufadienolide-like glycoside, scilliroside. The scilliroside content of the bulbs depends on the plant variety and the time of harvest but is typically below 0.5% of dry weight (Gentry et al., 1987; Verbiscar et al., 1986; see also Iizuka et al., 2001). The structure of scilliroside (Figure 3.14) was established in 1959 (Stoll, 1954; von Wartburg and Renz, 1959). (b) Identity, Physicochemical Properties, and Uses IUPAC name: (3)-(-d-glucopyranosyloxy)-17-(2-oxo2H-pyran-5-yl)-14-androst-4-ene-6,8,14-triol 6-acetate. Chemical Abstract name: (3,6)-6-(acetyloxy)-3-(-dglucopyranosyloxy)-8,14-dihydroxy-bufa-4,20,22-trienolide. CAS Registry Number: [507-60-8]. Empirical formula: C32H44O12; molecular weight: 620.7. Physicochemical properties  Scilliroside crystallizes from aqueous methanol as a hemihydrate with a melting point of 168–170°C. The compound is sparingly soluble in water, acetone, chloroform, and ethyl acetate but soluble in alcohols and acetic acid. Scilliroside is levorotatory: [ ]20 D  59.1 (c  1.0 in methanol).

193

Dried red squill powder loses its scilliroside content upon storage and when exposed to the atmosphere (Verbiscar et al., 1986). History and uses  The earliest record on the medical application of squill is in the Ebers Papyrus (Stoll, 1954; see also Aliotta et al., 2004). Herbalists recommended squill to treat dropsy and “to provoke urine, and open the stoppings of the liver and spleene.” The oldest among current rodenticides, red squill has been used since the 13th century (Stoll and Renz, 1942; Wax, 1995). It became more widely adopted in this century after the toxicant content of commercial products was biologically standardized. There are several Urginea (Charybdis) subspecies, and red and white squill varieties are distinguished. White squill preparations are the ones commonly used in human and veterinary (homeo)pathic medicine for its digitalis-like action (see, for example, European Agency for the Evaluation of Medicinal Products, 1999); it is only slightly toxic to rats. Red squill and scilliroside-based rodenticides are single-dose toxicants but their poor palatability seriously limits their use. Because scilla glycosides are very bitter, rodents consuming sublethal doses quickly learn to avoid the baits. Unlike scilliroside, however, its aglycone, scillirosidin [507-59-5], is tasteless but retains toxicity (oral LD50  0.35 mg/kg for female rats) (Rothlin and Schalch, 1952). Red squill formulations modified by Lactobacillus acidophilus -glycosidases did show improved acceptability but only females ate enough bait for a lethal dose (Verbiscar et al., 1989). Formulations  Technical-grade red squill powder can contain up to 28% scilliroside, and baits usually contain 0.01– 0.7% of the toxicant (Gratz, 1973). (c) Biological Properties Mode of action  Red squill exerts its action in three different ways: It causes emesis by local gastric action, affects the cardiovascular and the central nervous systems. Because rodents are unable to vomit, the selective toxicity of red squill is due to its quick and potent emetic action in humans and most nontarget animals that can regurgitate any ingested material. Although scilliroside may have cardiovascular activity per se, it is unlikely to pass the blood–brain barrier to cause convulsions, paralysis, and death. It has been suggested that the active toxicant is scillirosidin produced by -glycosidase-mediated metabolism from the parent glycoside. A further hydrolytic metabolite is desacetylscillirosidin, which is less toxic than the natural product. Thus, the large variations in toxicity and response time for red squill powders or scilliroside to different species, and even to individuals of the same species and sex, may be explained by differences in gut microflora responsible for the

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activation – and deactivation – of the ingested toxicant (Verbiscar et al., 1986). The insecticidal activity of U. maritima bulbs has been shown to be due to scilla glycosides, scillirosidin being the most toxic (Pascual-Villalobos, 2002). Leaf extracts, containing l-azetidine-2-carboxylic acid as the main bioactive component, are also insecticidal (see Civelek and Weintraub, 2004). (d)  Toxicity to Animals Acute toxicity  Red squill is considered to be a relatively safe rodenticide because animals other than rodents do not readily eat it. Typical LD50 values usually recorded at 2–5 days after administration of the toxicant to various laboratory and wild vertebrates are shown in Table 3.16. The oral toxicity of red squill concentrate to several nontarget farm animals was examined by Barnett et al. (1949). The unpalatable preparation was relatively nontoxic to pigs, cats, and dogs at 16 mg/kg, the maximum dose tested. Cats and dogs vomited shortly after treatment. Fowl were much less susceptible although fatalities did occur at doses higher than 400 mg/kg. Female rats are more sensitive than males to scilliroside (Dybing et al., 1952; see also Crabtree et al., 1939). This is similar to that observed for strychnine. The toxicity of red squill depends on altitude; for example, male rats were three times as resistant at 218 m as they were at 4328 m (Ward et al., 1940). Studies with baits containing the tasteless aglycone, scillirosidin, indicated that the bitterness of scilliroside is not the only reason for the learned bait avoidance (Verbiscar et al., 1989). The less polar aglycone crosses the blood–brain barrier and acts faster than the parent glycoside, causing rats to stop feeding once they experience the toxic effects (illness-based aversion learning). Poisoning syndromes  The typical course of scilliroside poisoning in rats after intravenous injection of an LD50 dose was as follows (Gold et al., 1947): Marked, but transitory weakness and ataxia occurred immediately after the injection. After a symptomless interval of 2–3 h, the second phase of poisoning, characterized by signs similar to curarization, began with signs of muscular weakness, ataxia, and the limbs sprawling out. Agitation appeared, and the limbs and the head were trembling. Soon hyperexcitability, then convulsions, sometimes resembling strychnine poisoning, developed. At this stage, the heartbeat was strong and rapid. The periods of convulsion alternated with marked generalized muscular depression. After 2–3 h in this stage, a sudden convulsion was followed by respiratory paralysis. Immediate dissection showed the heart still beating at slow rates but there were no pathological changes. Smaller doses produced muscular weakness, hyperexcitability, tremulous movements but no convulsions. At higher doses, the symptoms were different: The injection was followed by immediate

Hayes’ Handbook of Pesticide Toxicology

extreme prostration with labored respiration, feeble and irregular heartbeat, and death with cessation of heartbeat without convulsions within minutes. When the chest was opened immediately, the heart was found in standstill or in ventricular fibrillation, indistinguishable from cardiac death produced by digitalis glycosides. In another study with rats and mice, the effects of oral or intravenous (sub)lethal scilliroside and its aglycone were characterized by a latency period of several hours, emprosthotonos, hyperreflexia, rolling convulsions, and hypothermia. The manifestations as well as the lethal effect of scilliroside could be prevented or inhibited by prior oral administration of the anticonvulsant mephenytoin (Rothlin and Schalch, 1952). Rats poisoned orally with an LD50 dose showed initial lethargy, then spasmodic convulsions. Death was due to convulsions lasting for days and autopsy consistently revealed pulmonary hemorrhages (Brooks and Htun, 1980). Treatment  Barbiturates, magnesium sulfate, or chloral hydrate prevented convulsions but not death in mice. Sodium silicofluoride, citric acid, and especially oxalic acid, however, markedly antagonized the lethal effect of red squill in rats (Dybing et al., 1952). (e)  Toxicity to Humans Red squill preparations are emetic and that provides certain safety in humans. The poisoning symptoms produced by ingestion include abdominal pain and vomiting, cardiac irregularities, and convulsions. Accidental inhalation of red squill powder caused headache, vomiting, and diarrhea within 10 h, followed by lethargy and loss of appetite. No prolonged effects were observed (Barnett et al., 1949). Poisoning incidents  An apparent suicidal case of a man ingesting four tablets of a rodenticide containing a total of 12 mg scilliroside was described by Azoyan et al. (1991). Vomiting appeared a few minutes after ingestion and lasted for 2 days. An electrocardiogram, recorded 20 h after poisoning, revealed a complete atrioventricular block, which disappeared only on the fourth day. Blood pressure was 150/100 mmHg, and pulse varied between 80 and 100 beats/min. On the third day, digitoxin radioimmunoassay showed 11 nmol/l of scilliroside or its metabolite. In spite of treatment that consisted of gastric lavage, repetitive administration of activated charcoal, and maintenance of proper ion balance, the digitalization-like cardiac abnormalities disappeared only on the 12th day. A report of fatal squill poisoning of a hypothyroidic female who ingested two cooked bulbs of the plant as a folk remedy for arthritic pains was described by Tuncok et al. (1995). The symptoms resembled those of cardiac glycoside intoxication and included vomiting, seizures, hyperkalemia, atrioventricular heart block, and ventricular arrhythmia. The serum level of digoxin-like compounds,

Chapter | 3  Pest Control Agents from Natural Products

195

Table 3.16  Representative Acute Toxicity Data of Scilliroside or Other Squill Preparationsa Species, sex

Route

LD50 (mg/kg)

Note

References

Rat, male

oral

25.2

Red squill bait

Barnett et al. (1949)

Rat, female

oral

5.0

Red squill bait

Rat, male

oral

446

Rat, female

oral

165

Rat, male

oral

2.15

Rat, female

oral

0.43

Barnett et al. (1949) b

Verbiscar et al. (1986)

b

Verbiscar et al. (1986)

Red squill powder Red squill powder

Rothlin and Schalch (1952) Rothlin and Schalch (1952) c

d

Rat, female

oral

14–25

Rat, male

oral

5.3

Rat, female

oral

1.4

Bandicoot rat, male

oral

0.80

Bandicota bengalensis

Brooks and Htun (1980)

Bandicoot rat, female

oral

0.52

Bandicota bengalensis

Brooks and Htun (1980)

Rat, male

iv

2.05

Rothlin and Schalch (1952)

Rat, female

iv

0.50

Rothlin and Schalch (1952)

Rat, female

iv

1.25

Impure scilliroside

Gold et al. (1947)

Rat, female

iv

~12

“White squill”

Gold et al. (1947)

Mouse, male

oral

0.35

Rothlin and Schalch (1952)

Mouse, female

oral

0.43

Rothlin and Schalch (1952)

Mouse, male

oral

0.44

Dybing et al. (1952)

Mouse, male

sc

0.47

Mouse Mouse, male

oral oral

0.17 4.40

Rattus rattus

Barnett et al. (1949) Verbiscar et al. (1986) Verbiscar et al. (1986)

Dybing et al. (1952) d

Gray house mouse

Rothlin and Schalch (1952)

d

Rothlin and Schalch (1952)

d

Rothlin and Schalch (1952)

Field mouse

Mouse, female

oral

4.20

Field mouse

Guinea pig, male

oral

1.28

Rothlin and Schalch (1952)

Guinea pig, female

oral

1.00

Rothlin and Schalch (1952)

Rabbit, male

oral

7.70

Rothlin and Schalch (1952)

Rabbit, female

oral

7.50

Rothlin and Schalch (1952)

Cat

oral

6.00

Rothlin and Schalch (1952)

Cat

iv

0.120

Stoll and Renz (1950)

Cat

iv

0.133

Impure scilliroside

Gold et al. (1947)

Cat

iv

0.215

“White squill”

Gold et al. (1947)

Chicken

oral

20–25

Rothlin and Schalch (1952)

a

Data refer to pure scilliroside and laboratory, usually albino, rodent strains unless otherwise noted. From bulbs of clone #871 containing 0.10% scilliroside. 25% and 60% mortality at doses of 14.0 and 25.0 mg/kg, respectively. d Test was carried out with captured wild animals. b c

measured by immunoassay after admission, was 1.59 ng/l. Nine hours after ingestion, the patient lost consciousness, bradycardia was unresponsive to atropine, and a temporary cardiac pacemaker was applied that produced temporary clinical improvement. Trimethobenzamide was given for

persistent vomiting. The ventricular tachycardia, which occurred several hours later, failed to respond to lidocaine, and the patient died 30 h after ingestion. Polat et al. (2007) have reported an unusual, nonallergic contact dermatitis caused by a red squill preparation that

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was used to treat arthrodynia of joints (knee and wrist). The acute erythematous lesions were successfully treated with topical antibacterial cream and systemic antihistamines. Treatments  Treatment is symptomatic, involves gut decontamination (gastric lavage and administration of charcoal), and is the same as for digitalis overdose. Evaluation of a commercially available specific antidigitoxin antidote Fab (fragment antigen binding) fragment preparation in vitro showed a 4% cross-reactivity to and a 2.6  108 M1 apparent affinity constant for scilliroside, suggesting that this or related preparations could be useful antidotes for scilliroside poisoning (Sabouraud et al., 1990; see also Lapostolle et al., 2008).

3.5.3  Ricin (a)  Introduction The shrublike ornamental castor plant, Ricinus communis (Euphorbiaceae), is thought to be indigenous to northeast Africa and Asia but is now distributed worldwide. The seeds of the plant have three important constituents: an oil, castor oil, which is the glyceride of ricinoleic acid; a mildly toxic alkaloid ricinine [524-40-3]; and several isoforms of a highly poisonous glycoprotein ricin present up to 5% in the seeds. The yearly production of castor oil seed exceeds over 1 million metric tonnes (Food and Agricultural Organization of the United Nations, 2009). Purified castor oil has numerous uses. For example, it is an essentially harmless laxative, an industrial raw material, and an engine lubricant. It can also been converted by transesterification into biodiesel fuel. The toxic and proteinaceous properties of ricin were first established in 1887 by Stillmark at the university in Dorpat (now Tartu, Estonia), while the recognition of the specific binding properties of ricin by Ehrlich laid the foundation of immunology. The glycoprotein was obtained in pure, crystalline form in 1958 (reviewed by Balint, 1974; Olsnes, 2004). Ricin is one of the most toxic substances known and as such was once considered a chemical weapon (Agent W) but is now subject to the Chemical Weapons Convention. However, ricin has been used as a poison for criminal and terrorist purposes. Thus, the development of rapid, highly sensitive chemical and biological/ immunological analytical methods capable of detecting the toxin at or below ng/ml level is imperative (see, for example, Kalb and Barr, 2009; Lubelli et al., 2006; Uzawa et al., 2008). In cases where castor beans or crude ricin preparations are the suspected poisons, ricinine can also serve as a biomarker (see, for example, Mouser et al., 2007). The seeds of the leguminous jequirity plant (crab’s eyes or rosary pea, Abrus precatorius) contain abrin [1393-62-0], a related glycoprotein with biological properties similar to those of ricin.

Hayes’ Handbook of Pesticide Toxicology

(b)  Identity and Physicochemical Properties The CAS Registry Number of ricin is [9009-86-3]. In a pure state, ricin is a white crystalline powder. It is a water-soluble glycoprotein consisting of two polypeptides, termed A and B chains, which are linked by a disulfide bond (ASSB). The amino acid sequence of ricin (or ricin D as the toxic fraction from the beans is called) was determined by Funatsu et al. (1978, 1979). The A chain contains 265 amino acids and has a molecular weight of 32 kDa; its sugar content is 2.6%. The isoelectric point of the A chain is 7.34. The B chain of 260 amino acids and four internal disulfide bonds has a molecular weight of 34 kDa and its sugar content is 6.4%. The A chain has enzymatic properties (ribosomal RNA N-glycosidase, EC 3.2.2.22) responsible for the toxicity of ricin, while the B chain is a lectin binding to galactose-containing glycoproteins and glycolipids on the surface of target cell components. The threedimensional structure of ricin has been solved by X-ray crystallography studies (reviewed by Lord et al., 1994). The physicochemical and photochemical properties of ricin have recently been reviewed (Gaigalas et al., 2007). Stability  Because the chaff left over from castor oil processing can be used as animal fodder, a great deal of effort was devoted to its detoxification (Balint, 1974; European Food Safety Authority, 2008). High-temperature denaturing (80°C for 1 h) and chemical methods (oxidation with potassium permanganate, hydrogen peroxide, iodine, etc.) were recommended to destroy the toxin (see, for example, Barnes et al., 2009). In the presence of 2-mercaptoethanol, which reduces the disulfide bond joining the A and B chains, the toxicity of ricin is lost; removal of 2-mercaptoethanol, however, allows the reconstitution and reactivation of the toxin. Ricin is degraded by papain but only slowly by trypsin. The fate of ricin in the body is incompletely understood. Use  The toxin-rich chaff byproduct of castor oil manufacture has been used to kill mice and moles. Conjugates of ricin and cell-specific antibodies are experimental anticancer immunochemotherapeutic agents (Sandvig and van Deurs, 2005; Stirpe et al., 1992). Recently, transgenic rice and maize engineered to produce a fusion protein comprising the Cry1Ac endotoxin of Bt and the ricin B lectin subunit have been shown to be insecticidal to insects that are otherwise tolerant to Cry toxins (Mehlo et al., 2005). (c)  Biological Properties Mode of action  Once it was thought that the toxic action of ricin preparations in mammals is due to its hemagglutinating effect, but this activity was shown to be associated with the structurally similar, but nontoxic agglutinins present in the castor bean. It is now well established that ricin inhibits protein synthesis in eukaryotic systems by catalytically

Chapter | 3  Pest Control Agents from Natural Products

inactivating the 60S subunit involved in the translation process. The structural aspects of the mode of action of ricin and other ribosome-inactivating protein (RIP) toxins from plants and fungi were thoroughly reviewed (Kozlov et al., 2006; Stirpe and Battelli, 2006). Briefly, the B chain binds to galactose/N-acetylgalactosamine-containing glycoproteins and glycolipids in eukaryotic cells. The binding to surface receptors can be inhibited by galactose or lactose in vitro. It appears that both chains facilitate the penetration by endocytosis of the toxin into the cell. The B chain, however, aids the toxin in translocating to endosomal targets as well. Once in the cytosol, the A chain cleaves a single adenine base from the 28S ribosomal RNA within the 60S ribosomal subunit, rendering it unable to bind the elongation factor 2 which consequently leads to an arrest of protein synthesis. A single A chain molecule can inactivate 1500 ribosomes per minute and kill the cell. In addition to inhibiting protein synthesis, ricin was shown to induce apoptosis, cause oxidative stress, release proinflammatory cytokines, alter cell membrane structure and function, and damage nuclear DNA (reviewed by Stirpe and Battelli, 2006). Lipase activity of ricin was also demonstrated (Morlon-Guyot et al., 2003). 125

Distribution, metabolism, and excretion  I-Labeled ricin injected either intravenously or intraperitoneally was distributed in various tissues, accumulating in the spleen, kidneys, heart, liver, and thymus (Fodstad et al., 1976). Urinary excretion of radioactive degradates, but not intact ricin, peaked 5–7 h after injection and was complete within 10–20 h. Because the intact toxin is resistant to proteolytic enzymes in vitro but the separated chains are considerably more vulnerable, it has been suggested that degradation in tissues occurs after reduction of the ASSB disulfide bridge and separation of the chains. In a nasal-inhalation study with mice, [125I]ricin was initially accumulated in the lungs with an approximately 40 h half-life but quickly distributed into the trachea (especially the associated thyroid) and the gastrointestinal tract; little evidence of systematic dissemination of the toxin was observed at this inhaled dose (Doebler et al., 1995). The size of the aerosolized particles affects the distribution and toxicity of the toxin: inhalational exposure of mice to ricin particles of 1 m (at 4.5  LD50 dose) resulted in 100% mortality by 72 h, while for toxin particles 5 m (at 3.7    LD50 dose determined of the 1-m particle) no mortality was observed (Roy et al., 2003; see also Griffiths et al., 2007). This study also showed that the 1-m particles accumulated mainly in the lungs (60%), while larger particles deposited higher in the airways, that is in the trachea (80%). Comparing different mice strains, Godal et al. (1984) found that more sensitive strains concentrated higher amounts of [125I]ricin in their liver, spleen, and kidneys. The liver was rich in modified ricin and also in dissociated and modified A chains. Considerable amounts of the

197

toxin accumulated in the adrenal cortex and bone marrow as well. In a subsequent phase I study with human cancer patients, the largest intravenous ricin dose with tolerable side effects (nausea and muscular pain) was 18 g/m2 with initial plasma concentration of about 25 ng/ml; concomitant anti-ricin antibody formation was also demonstrated. Cook et al. (2006) have used an amplified ELISA with a limit of sensitivity of about 200 pg/ml to compare the tissue distribution of sublethal doses of crude ricin toxin (50% ricin content) following pulmonary and oral dosing to rats (250 g average weight). After pulmonary instillation of 0.8 g crude ricin, the total ricin content of the lung increased from 11.4 ng at 24 h to 40.9 ng at 48 h (corresponding to 7.12–11.38 ng/g tissue). However, the amount of ricin recovered from this tissue was 1–5% of the original dose. Alternatively, 24 h after an oral dose of 2 mg crude ricin per rat (8 mg/kg), the toxin was accumulated in the liver, spleen, gastrointestinal tract, and kidney (corresponding to 9.5, 31.7, 39.4, and 10.9 ng/g tissue, respectively). In this case, the total recovered amount was below 1% of the administered dose. It is also noteworthy, that blood, typically used for diagnostic purposes, contained only 1.4 ng/ml ricin. (d)  Toxicity to Laboratory Animals Acute toxicity  Although the mode of action of ricin at the molecular level is known, the mechanisms responsible for the clinical and lethal effects of the toxin are still inadequately understood. Representative animal toxicity data for ricin administered by different routes are shown in Table 3.17. The variations in the acute toxicity values reported in the literature are mainly due to purity differences of the preparations used in the tests (reviewed by Balint, 1974). The symptoms of ricin poisoning manifest slowly, usually 12 h after administration, and include rather sudden outbursts of convulsions and opisthotonos, followed by paralysis of the respiratory center, eventually leading to death. Accidental poisoning is commonly due to ingestion of castor beans (for recent examples, see Aslani et al., 2007; Soto-Blanco et al., 2002). Laboratory tests with seeds showed hen to be the most resistant species (the lethal dose was 14 g/kg); sheep and horse were more sensitive (lethal doses were 1.25 and 0.10 g/kg, respectively). The toxin is pyrogenic in mammals (Balint, 1993). In the serum of animals treated with ricin, antibodies specific to ricin have repeatedly been detected (see, for example, Griffiths et al., 2007). Ricin is highly toxic upon injection and inhalation. Using transmission electron microscopy, Brown and White (1997) (see also Griffiths et al., 1995b, 2007) examined the histopathological changes in the lungs of rats upon ricin inhalation. The animals were exposed to an LCt30 (the concentration in air that killed 30% of the exposed animals) of 11.21 mg/min/m3 dose of the toxin. Necrotic changes were evident in the capillary endothelium and type I epithelial cells, accompanied by intraalveolar edema 12–15 h after exposure.

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Table 3.17  Acute Toxicity of Ricin Species, sex

Route

LD50 (g/kg)

Rat

oral

20,000

Minimal lethal dose (g/kg)

References Cook et al. (2006)

Rat

ip

38.2

Rat

inhalation

3.7a

50

Rat

inhalation

9.8

b

Mouse

ip

37.0

Mouse

ip

2.6

Olsnes and Pihl (1973)

Eperjessy et al. (1965) Griffiths et al. (1995b) Griffiths et al. (1995b)

50

Eperjessy et al. (1965)

Mouse

ip

4.0

Olsnes et al. (1974)

Mouse

ip

8.0c

Olsnes et al. (1974)

Mouse, female

ip

Mouse, male

iv

Mouse

iv

Mouse

sc

25 2.5

Muldoon and Stohs (1994) Fodstad et al. (1976)

2.7 24

Fodstad et al. (1979) Poli et al. (2007)

Mouse

oral

20,000

Poli et al. (2007)

Mouse, various strains

inhalation

2.8–11.2

Poli et al. (2007)

Guinea pig

ip

37.8

50

Eperjessy et al. (1965)

Rabbit

ip

7.82

10

Eperjessy et al. (1965)

Cat

ip

29.0

40

Eperjessy et al. (1965)

Dog

iv

African vervet monkey

ip

30.3

Balint (1993)

African vervet monkey

inhalation

5.8

Wannemacher and Anderson (2006)

Rhesus monkey

inhalation

Rhesus monkey

inhalation

1.6–1.75

36.5 15

Fodstad et al. (1979)

Wilhelmsen and Pitt (1996) Wannemacher and Anderson (2006)

a

Ricin from R. communis var. “Hale Queen”. Ricin from R. communis var. zanzibarensis. Ricin reconstituted from A and B chains.

b c

Wong et al. (2007) have recently demonstrated that tracheal instillation of sublethal dosages of ricin in mice induces localized inflammation in the lungs with minimal evidence of systemic effects, which is in agreement with earlier distribution studies. A lethal dose of the toxin, however, causes not only severe hemorrhagic inflammatory response in the pulmonary system but enters the vascular system and initiates inflammation in multiple organ sites, especially in the kidneys. Wilhelmsen and Pitt (1996) studied in rhesus monkeys the inhalational toxicity of 21–41.8 g/kg doses of aerosolized ricin. The major intoxication symptom was a rapid onset of dyspnea occurring after a 20–24 h lag period after exposure. Doses above 36.5 g/kg caused death 36–48 h after exposure. In addition to skin elasticity indicating dehydration, autopsy findings confined to the exposed organ, that is the respiratory tract, thus differed from those

reported for parenteral ricin intoxication in other species. They included clear foamy fluid in the trachea and mainstream bronchi; serous fluid in the thoracic cavity; diffusely wet, mottled red lungs with multifocal necrosis; acute inflammation of airways, and lesions along the pulmonary lymphatic drainage course. In some monkeys bilateral adrenocortical necrosis was also seen. The cause of death was attributed to asphyxiation following massive pulmonary alveolar flooding. Treatment  Treatment is entirely symptomatic. No practically useful antidote exists, but immunization could protect animals from the lethal effect of inhaled ricin (Griffiths et al., 1995a). Muldoon and Stohs (1994) found that dexamethason and difluoromethylornithine significantly extended the survival time of mice treated with an intraperitoneal LD100 dose (25 g/kg) of ricin, but death could

Chapter | 3  Pest Control Agents from Natural Products

not be prevented. Mabley et al. (2009) have recently demonstrated that subcutaneous administration of 0.4 g/kg nico­tine, a known activator of cholinergic anti-inflammatory pathways, delayed mortality and reduced systemic organ failure in mice previously injected with lethal doses of ricin (50 or 100 g/kg, intraperitoneally). When lactose, a known inhibitor of ricin binding to cell surface receptors in vitro, was administered together with ricin, the organ distribution changed: Almost 80% of the toxin injected was found in the liver after 30 min, compared to 48% without lactose, and the amount in the other organs was reduced. Moreover, on coinjection, 0.3 ml of 0.25 M lactose protected mice from a 0.25-g lethal dose of ricin (Fodstad et al., 1976). (e)  Toxicity to Humans As with animals, the toxic effects of ricin have a latency period of several hours after administration. The initial clinical symptoms of ricin poisoning are nonspecific and include general malaise, nausea, violent vomiting, bloody diarrhea and tenesmus, thirst, dilation of the pupils, conjunctivitis, shivering, and fever. Additional features of poisoning are mucosal damage, gastrointestinal bleeding, edema, renal failure, and vascular collapse. In severe poisoning, convulsions may precede death. According to experiments with nonhuman primates and a small number of human nonfatal poisoning cases (Associated Press, 2008; Poli et al., 2007), symptoms of inhalational exposure are respiratory distress, cough, fatigue, fever, and arthralgias that could be followed by inflammation of airways, pneumonia, kidney failure, and coma. Based on animal data and human poisoning incidents, the lethal dose is estimated to be 1–20 mg/kg (3–10 castor beans, depending on size, moisture content as well as on the degree of mastication) by the oral route and 5–10 g/kg by injection or inhalation. For a general discussion on ricin toxicology and poisoning as relevant to humans, see the review by Audi et al. (2005). The well-known allergy to castor bean and derived industrial products is due to proteinaceous allergens and not to ricin. Szalai et al. (2005), however, have described a series of allergic reactions among researchers working with ribosome RIPs, including ricin. Poisoning incidents  With a few exceptions, all reported human poisoning cases were due to ingestion of seeds of the castor plant, although death was exceptional (Audi et al., 2005; Rauber and Heard, 1985). Lüfti (1935) described a fatal castor bean poisoning case. Two hours after ingesting 15–20 seeds, a man started to have nausea and abdominal pains, vomited, and later enduring diarrhea set in. The patient died as a result of complications of circulatory disorders, nephritis, and uremia 13 days after ingestion. Autopsy revealed that the kidneys, liver, heart, and spleen were hemorrhagic, necrotic, and inflamed.

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Children aged 7–12 ingesting one or two castor beans received treatment consisting of ipecac-induced emesis, charcoal, and, to prevent hemolysis, alkalinization of urine (Rauber and Heard, 1985). Except for mild diarrhea, no symptoms of castor bean toxicity were noticed. The same authors reported a suicide attempt of a patient ingesting at least 24 chopped castor beans. Following treatment consisting of induced emesis and gastroscopic removal of the bean particles, the patient was asymptomatic. Another recent case involved a young adult who ingested about a dozen castor beans, some of them chewed (Aplin and Eliseo, 1997). Four hours after ingestion, the patient started to feel ill, had severe abdominal pains, and vomited. The patient was given intravenous fluid, antiemetics, and charcoal 6 h after ingestion. After a mild hypokalemia (2.8 g/ml) on the second day, recovery was complete by the third day. Kopferschmitt et al. (1983) described a case of a 21-year-old man ingesting 30 castor beans (some masticated) with suicidal intent. Symptoms typical of ricin poisoning were observed. Radioimmunological analysis for ricin showed first-day plasma level of 1.5 ng/ml; the toxin appeared in the urine only on the third day (0.3 ng/ml). Treatment involved saline and glucose infusions and the patient recovered. A more recent suicidal case involved a young patient who subcutaneously injected an extract of castor beans (Targosz et al., 2002). The main symptoms presented 36 h after the injection were fatigue, nausea, dizziness, compression of chest, abdominal pain and muscular pain of extremities with numbness, anuria, tachycardia, hypotension, and metabolic acidosis. At the site of the needle insertion, suggillation and edema were seen. Bloody diarrhea and hemorrhagic diathesis as well as liver, kidney, cardiovascular and respiratory systems failure also developed. In spite of symptomatic intensive care, fatal asystolic arrest ensued 44 h after the injection of the poison. The most publicized fatal ricin poisoning case involved the politically motivated murder of the Bulgarian playwright Georgi Markov (Crompton and Gall, 1980). Ricin, which actually could not be directly identified, was dispensed in a small, hollow metallic sphere and probably shot from a modified umbrella into the right thigh of the victim. The amount of ricin in the sphere could be no more than 1 mg. Within a few hours, the victim was in pain, and the wound became inflamed. The following day, he had a high fever and vomited. Later, his blood pressure and temperature fell, the pulse rate rose to 160 beats/min, the white cell count was 26,300/ml, and renal tubular necrosis set in. Death was due to cardiac arrest on the third day after the injury. Autopsy revealed interstitial hemorrhages in the intestines, testicles, pancreas, and inguinal glands. Microscopy showed myocardial hemorrhages. Treatment  Because there is no demonstrated antidote to ricin poisoning in humans, treatment is symptomatic and

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supportive with special attention to vasopressor therapy, fluid and electrolyte management, and avoidance of shock. Gastrointestinal decontamination can involve charcoal in nonvomiting patients and gastric lavage could be considered for patients presenting within 1 h of ingestion. For inhalational exposure, supportive treatment to counteract acute pulmonary edema and respiratory distress is indicated. A recently developed recombinant ricin vaccine appears to be promising (Vitetta et al., 2006). Effective anti-ricin antibodies have also been reported (see, for example, Wang et al., 2007).

3.5.4  Salmonella Bacteria (a)  Identity Names: Salmonella enteritidis and S. typhimurium. (b)  History The use of Salmonella as a rodenticide, especially against field mice, was recommended soon after the bacterium was isolated, characterized, and evaluated in the 1890s. The effectiveness of these products against wild rodents, however, was found to be questionable; moreover, the extensive use of Salmonella-based rodenticides in the first half of the 20th century increased the prevalence of infection carriage in rodents and proved to be a hazard to human health. Salmonella-containing rodenticides were banned in the United States in the 1920s and, in the early 1960s, in the UK, as well as in many other countries (Wodzicki, 1973). Although it would appear that S. enteritidis pathogens for rodent control have only historical significance, this type of product resurfaced in some countries in the 1990s. (c)  Human Poisoning Incidents Acute enteritic infections are usually without fatalities, but there have been cases that could be traced to Salmonellarodenticide use causing deaths. One of the first reported epidemics was caused by the “Liverpool virus” rat poison placed in the homes of patients (Handson et al., 1908). Twelve persons who were infected by the bacterium showed symptoms of vertigo and abdominal cramps developing within an hour or two, vomiting, and diarrhea. The patients had high fever, were thirsty, and urine was entirely suppressed for over 12 h. Recovery was within 10 days. A similar infection was reported by Dathan et al. (1947). In both cases, the pathogen was S. enteritidis but apparently different variants. In the UK between 1944 and 1955, there were 23 fatalities of the 1680 reported human infection cases of which 75% were caused by S. enteritidis var. jena and the rest by S. enteritidis var. danysz (Taylor, 1956). Recently, the appearance of a salmocoumarin (“Biorat” from Cuba) rodenticide containing 0.02% warfarin and 25% S. enterica subspecies enterica, bioserotype enteritidis

(subgroup I), Group D var. 7F4, was reported (Friedman et al., 1996; Threlfall et al., 1996). (Note: “Biorat” was advertised on internet sites as recently as August 2009.) A 2001 sample of such a formulation was shown to contain 0.002% “hydroxycoumarin”, 200,000 CFU/gram of S. enteritidis phage type 6a and rice (Painter et al., 2004).

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Chapter | 3  Pest Control Agents from Natural Products

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Chapter 4

Public Health Pesticides Robert J. Novak1 and Richard L. Lampman2 1 2

University of Alabama at Birmingham, Birmingham, Alabama Illinois Natural History Survey, Champaign, Illinois

4.1  Introduction The spectacular success with synthetic insecticides in the decade immediately following World War II generated widespread enthusiasm that the major scourges of mankind, such as yellow fever, malaria, and typhus, could be conquered. From 1955 to 1969, the World Health Organization (WHO) Global Malaria Eradication Program eliminated malaria from 36 countries, primarily by spraying inside human habitats with a relatively inexpensive residual insecticide, DDT (Metcalf and Novak, 1994). The annual number of malaria cases in India was reduced from 75 million to 150,000 and deaths from 750,000 to 1500 during the period from 1952 to 1966, largely due to the use of organochlorine insecticides (Metcalf, 1998). Unfortunately, by the late 1960s, governments throughout the world were forced to reevaluate vector and pest control techniques. In addition to insecticide resistance, there was a precipitous decline in beneficial insect species, outbreaks of secondary pests, contamination of the environment and foodstuffs, and bioaccumulation of pesticide residues in nontarget organisms, including humans (Brogdon and McAllister, 1998; Brown et al., 1976; National Academy of Sciences, 1980; WHO, 1972, 1992). Furthermore, once malaria seemed to be under control, there was a loss of trained vector control specialists in some countries, and in other countries pathogens started to develop resistance to antimalarials (Gubler, 1996, 1998). The escalating problems with agricultural pests and public health vectors sparked a conceptual change in insect control, which shifted from an almost exclusive reliance on insecticides to a blend of cultural, physical, biological, and chemical methods (Curtis, 2001; Luckmann and Metcalf, 1994; WHO, 1995). Thus, currently the goal of WHO is disease management through a combination of site-specific integrated pest management (IPM) techniques in coordination with medical diagnosis, treatment, and epidemiological trend analysis (see http://www.who.int/health-topics/malaria/en). In general, IPM is an applied ecological approach to curtailing the damage of injurious plants and animals to Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

economically acceptable levels using a variety of longterm and short-term interventions that are environmentally and socially sound. With the advent of IPM, it is no longer proper to discuss the use of pesticides without including the context in which they are to be used and their potential detrimental side effects (WHO, 1995). However, pesticides remain a critical component of vector management programs because they provide a rapid and highly efficacious means of reducing arthropod populations, particularly when the potential for disease transmission is high. In developing countries, approximately 10% of total pesticide use is for vector management (Leng, 1999). In the United States, it has been estimated that $150 million is spent on vector control and surveillance for arboviral encephalitides (see http://www.cdc.gov/ncidod/dvbid/Arbor/arbofact.htm). To provide a general review of public health pesticides, we briefly cover the following topics: (1) common terms and concepts in medical entomology in order to introduce the diverse relationships among humans, arthropods, and disease pathogens; (2) the historical and current impact of arthropods on human health; (3) the basic concepts of vector management; (4) a brief list of noninsecticidal alternatives; (5) the classification schemes and general properties of public health pesticides and repellents; and (6) specific examples of pesticides in vector management, particularly mosquito control. Our approach to vector management is analogous to that of modern medicine. Vector control specialists should have the same degree of understanding about insecticides and acaricides as a physician has for prescription drugs. The impact of these chemicals on wellness of a human body or an ecosystem has to be addressed before and after treatment. The prevailing doctrine should be “Do no harm.”

4.2  Definition of terms in   vector-borne diseases The science that deals with the impact of arthropods on humans is typically called medical entomology, despite 231

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the fact that noninsect species such as ticks, mites, and spiders are generally included. The interaction of humans and arthropod pests can be categorized into direct and indirect effects (Harwood and James, 1979). Direct effects are physical, physiological, or psychological responses of humans to the actions or the presence of arthropods. Direct effects include blood loss, ectoparasitic skin irritations (dermatitis), allergic reactions, envenomization, urticating hairs of caterpillars, invasion of tissues and organs as with endoparasites (such as myiasis with dipterans), and a wide range of mental disturbances that range from prevention of normal activities to delusional parasitosis (Table 4.1). Common arthropods that cause direct effects include cockroaches, bedbugs, blister beetles, spiders, scabies and chigger mites,

head and pubic lice, horse flies and deer flies, stable flies, mosquitoes, ticks, and various species of ants, bees, caterpillars, and wasps. Indirect effects are those where arthropods serve as carriers (vectors) of etiological agents (pathogens) to humans and animals (hosts). The pathogen transfer may be by mechanical transmission, carried on the outside of the arthropod, or by biological transmission, carried internally and typically transmitted through some arthropod product such as salivary secretions or feces. In biological transmission, the pathogen may undergo partial or complete development, as well as multiplication, within the vector. The principal vectors worldwide are flies and mosquitoes (Diptera), ticks (Acarina), fleas (Siphonaptera), sucking

Table 4.1  Examples of Arthropods by Order and Family That Cause Indirect Annoyances to Humans and Animals through Venoms and Toxins, Allergens, and Pests Blood loss—prevention of normal activities Culicidae (mosquitoes): Aedes vexans, Ae. sollicitans, Ae. dorsalis, Culex tarsalis, Psorophora columbiae Simuliidae (black flies): Simulium vittatum, S. venustrum, Prosimulium hirtipes Ceratopogonidae (no-see-ums): Culicoides canithorax, C. melleus, C. furens, C. obsoletus Muscidae: Stomoxys (stable flies): Stomoxys calcitrans, S. irritans Tabanidae (deer and horse flies): Chrysops sp., Tabanus sp. Cimicidae (bedbugs): Cimex lectularis Ixodidae (hard ticks): Dermacentor andersoni, D. variabilis, Rhipicephalis sanguineus Dermatosis/dermatitis (skin irritations) Pediculidae (head and pubic lice): Pediculis humanis capitus, Phthirus pubis Sarcoptidae (scabies mites): Sarcoptes scabiei, Sarcoptes sp. Trombiculidae (chigger mites): Trombicula autumnalis, T. alfreddugesi, T. splendens Occupational mites: Pyemotis tritici, Tyrophagusputrescentiae, Glycyphagus domesticus Liparidae (tussock moth) Meloidae (blister beetles) and Oedemeridae (false blister beetle) Invasion of tissues and organs Flies (myiasis), beetles, and other insects Hippoboscidae (louse flies): Melophagus ovinus Muscidae (house, horn, stable flies): Haematobia irritans Oestridae (bot flies): Gastrophilus equi, G. intestinalis, Hypoderma lineatum Calliphoridae (blue bottle flies): Cochliomyia hominivorax (screw worm) Envenomization, bites, and urticating hairs (wasps, ants, bees, spiders, caterpillars, stings, bites, localized reactions) Apidiae (honeybees, bumblebees): Apis mellifera, Apis mellifera scutellata Vespidae (wasp, yellow jackets, hornets): Vespula sp., Dolichovespula sp. Mutillidae (velvet ants): Ephuta, Photomorphus, Pseudomethoca, Sphaeropthalma, and Timulla sp. Formicidae (ants): Solenopsis sp. (fire ants), Camponotus sp. (carpenter ants), Formica sp. (wood ants) Lepidoptera (caterpillars or larvae, urticating hairs)

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Table 4.1  (Continued) Saturniidae: Hemileuca oliviae, Automeris sp. (Io moths) Lymantriidae: Nygmia phaeorrhoes (brown-tailed moth) Megalopygidae (flannel moths): Megalopyge lanata Arctiidae: Lithosa caniola, Arctia caja, Euchaetias egle Nymphalidae: Nymphalis antiopa Noctuidae: Acronicta lepusculina, A. oblinita, Catocala sp. Thysanoptera (thrips): Thrips tabaci, Chirothrips aculeatus, Heliothrips jumipennis, H. sudanensis, H. indicus Coleoptera (beetles): Meloidae (blister beetles): Lytta vesicatoria; Oedemeridae (false blister beetle) Injection of toxin Acarina: Metastigmata (can cause tick paralysis) Ixodidae (hard-bodied ticks): Amblyomma sp., Dermacentor sp., Hyalomma sp., Rhipicephalus sp., Boophilis sp. Argasidae (soft-bodied ticks): Ornithodorus sp. Araneida (spiders): ������Phoneutria nigriventer (banana spider), Loxoceles reclusa (brown recluse spider), Lactrodectans mactans (black widow spider) Chilopoda (centipedes): Scolopendra subspinipes Scorpiones (scorpions): Centruroides sp. Allergic reactions (localized and systemic), insect body parts, by-products, bites, anaphylactic shock Apidae (honeybees, bumblebees): Apis mellifera, A. mellifera scutellata Vespidae (wasp, yellow jackets, hornets): Vespula sp., Dolichovespula sp. Mutillidae (velvet ants): numerous species Formicidae (ants): Solenopsis sp. (fire ants) Culicidae (mosquitoes): numerous species Simuliidae (black flies): numerous species Acarina (house dust mites): Dermatophagoides pteronyssuss, D. farinae, numerous other species Blattidae (cockroaches): ingested feces, numerous species Acarina: Ixodidae and Argasidae (soft- and hard-bodied ticks) Entomophobia and delusionary parasitosis Numerous examples for all insects, spiders, ticks, and mites

lice (Anoplura), and true bugs (Hemiptera). The major pathogens include viruses, rickettsia, bacteria, helminths, and protozoa. Viruses transmitted by arthropods are called arboviruses. Zoonoses are diseases in which the pathogens are maintained in vertebrate hosts other than humans (i.e., yellow fever and arboviral encephalitis). Anthroponoses are those diseases in which humans are the only known vertebrate host (i.e., malaria and epidemic typhus). Common vectors and pathogens for several important tick- and mite-borne diseases are listed in Table 4.2. Insectborne pathogens and vectors are listed in Table 4.3, and those transmitted by mosquitoes are listed in Table 4.4. Mosquitoes deserve special recognition because they are responsible for transmitting more pathogens to humans,

causing greater hardships worldwide, than all the other vector-borne diseases combined (Busvine, 1993). Examples of vectors and associated diseases include Anopheles mosquito species and the different types of malaria; the human body louse, Pediculus humanus humanus L., and epidemic typhus; the mosquito, Aedes aegypti L., and yellow fever and dengue; the oriental rat flea, Xenopsylla cheopis (Rothschild), and plague; the tsetse flies, Glossina spp., and African trypanosomiasis; the Triatominae bugs and American trypanosomiasis; Simulium species and onchocerciasis; Leptotrombidium tick species and scrub typhus; and the black-legged tick or deer tick, Ixodes scapularis, and Lyme disease (Eldridge and Edman, 2000). Arthropods may be both nuisance pests (direct effects) and disease vectors

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Table 4.2  Vector-Borne Pathogens Transmitted by Mites and Ticks Disease

Pathogen

Vector Mites Bacteria

Rickettsial pox Scrub typhus

Rickettsia akari R. tsutsugamushi

Liponyssoides sanquineus Trombiculidae and Leptotrombium species Ticks

Arboviruses Tick-borne encephalitis Omsk hemorrhagic fever

Flavivirus Flavivirus

Kyasanur forest disease Powassan encephalitis Russian spring–summer fever

Flavivirus Flavivirus Flavivirus Haemphysalis concinna Bunyaviridae, Nairovirus Reoviridae and other tick species

Crimean–Congo hemorrhagic fever Colorado tick fever

Ixodes ricinus Dermacentor pictus D. marginalis Ixodes persulcatus Haemaphysalis species Dermacentor andersoni Ixodes persulcatus H. japonica douglasi Hyalomma marginatus and other tick genera Dermacentor andersoni

Bacteria Tick-borne relapsing fever Tularemia Rocky Mountain spotted fever

Borrelia recurrentis Francisella tularensis Rickettsia rickettsia

Lyme disease

Borrelia burgdorferi

Ehrlichiosis

Ehrlichia sp.

Ornithodorus spp. Ixodes and Dermacentor sp. Dermacentor andersoni D. variabilis Ixodes scapularis Dermacentor variabilis D. variabilis, Amblyomma americanum

Protozoa Texas cattle fever East coast fever Babesiosis

Babesia bigemina Theileria parva Babesia microti

(indirect effects). For example, the forest day (Asian tiger) mosquito, Aedes albopictus Skuse, which was introduced into the United States presumably in used tires from Japan, is a potential vector of several arboviruses; however, even in the absence of disease transmission, it is an aggressive daytime biter, making it a major pest wherever it becomes established (Hawley, 1988; Hawley et al., 1987; Mitchell, 1991, 1995; Moore and Mitchell, 1997; Novak, 1995; Shroyer, 1986). The transmission of a pathogen may be classified as either horizontal or vertical. Horizontal transmission is from an infected animal host via vector to a new, uninfected animal host. Some pathogens, particularly viruses, are also transmitted vertically. Vertical transmission may be between an infected female host and her offspring or between developmental stages of a vector arthropod. An infected female insect may transfer pathogens to the eggs (transovarial transmission) and the pathogens may be

Boophilus annulatus Rhipicephalus appendiculatus Ixodes sp.

transferred from the eggs to the different developmental stages (transtadial transmission). Adult males may also transmit pathogens to females during mating (venereal transmission). A key concept in medical entomology is that all species of arthropods are not capable of transmitting all possible pathogens. Evolutionarily, a pathogen has adapted by various means to the biological and ecological barriers associated with its invertebrate vectors and vertebrate hosts. Thus, mosquitoes, in general, do not transmit the human immunodeficiency virus, Culex species are not vectors of malaria, and Anopheles species are not a concern for arboviral encephalitis. For a pathogen to pass from an infected host to a blood-feeding vector, there must be a relatively high level of circulating pathogens in the host because of the extremely small percentage of total blood taken in the blood meal. Many disease agents, such as arboviruses and the malaria parasite, have overcome this hurdle by causing either a viremia or a periodic release of

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Table 4.3  Vector-Borne Pathogens Transmitted by Nonmosquito Insects Insect species

Disease

Pathogen

Distribution

Arboviruses Phlebotomus papatasi and other species Culicoides sp.

Sand fly fever

Bunyaviridae

Africa, Asia, Europe

Bluetongue virus

Reoviridae

Africa, Asia, United States

Bacteria Xenopsylla cheopis and other rodent fleas Xenopsylla cheopis, various species Pediculus humanus

Plague

Yersinia pestis

Worldwide

Murine typhus

Rickettsia typhi

North America, Europe

Epidemic typhus

Rickettsia prowazekii

Worldwide, scattered foci

Pediculus humanus Phlebotomus spp. Chrysops spp.

Louse-borne relapsing fever Bartonellosis Tularemia

Borrelia recurrentis Bartonella bacilliformis Francisella tularensis

Mainly Africa South America Worldwide

Protozoa Glossina spp. Triatominae spp.

African trypanosomiasis, African sleeping sickness American trypanosomiasis (Chagas disease)

Phlebotomus spp.

Visceral leishmaniasis

Phlebotomus spp.

Cutaneous leishmaniasis

Lutzomyia spp.

American leishmaniasis

Trypanosoma gambiense, T. rhodesiense Trypanosoma cruzi Panstrongylus spp. Rhodnius spp. Leishmania spp. L. donovani

Africa

Leishmania spp. L. tropica L. braziliensis Psychodopygus spp.

North Africa, Middle East

Central and South America Mediterranean, North Africa, Middle East, Asia, Central and South America

Central and South America

Nematodes Simulium spp. Chrysops spp.

Onchocerciasis Loiasis, eye worm

an infective stage in their hosts, respectively. For example, Plasmodium parasites must pass through several biological barriers in the mosquito vector, including the midgut and hemolymph, in order to eventually concentrate in internal organs such as the salivary glands where they are transmitted to a host with the next blood-feeding cycle (Beier, 1998). If the pathogen completes this passage and can infect a host, the arthropod carrier is called a competent vector, even if it has not been demonstrated to be an ecologically significant one. Transmission efficiency may vary considerably among species, and their role in transmission in an area may depend on their abundance, longevity, and feeding behavior. Animal hosts that do not produce a high level of circulating pathogens for a biologically significant period of time are dead-end hosts. They may contract the disease but are not important in maintaining the transmission cycle. Such is the relationship of humans and many arboviruses, such as St. Louis encephalitis virus (SLEV)

Onchocerca volvulus Loa loa

Central and South America, Africa West and central Africa

and eastern equine encephalitis virus (Monath, 1988). Transmission does not invariably lead to disease in hosts because many sylvatic hosts are asymptomatic. The pathogen also faces the possibility of being introduced by a competent vector into a host capable of clearing the invading pathogen or a previously infected host with a stimulated immune system. Often, vertebrate hosts become immune for an extended period of time after an initial infection; however, there are also cases of recrudescence (in which the host exhibits clinical symptoms of a disease from a previous infection after a prolonged period of recovery). Pathogens can also evade the immune response of hosts by changing the antigen signal presented to the host’s immune system (Mandell, 1990; Manson-Bahr and Bell, 1987). The time from the introduction of the pathogen into the host to the first clinical expression of the disease is called the intrinsic incubation period. The extrinsic incubation period is the time from which the vector acquires the pathogen

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Table 4.4  Major Vector-Borne Pathogens Transmitted by Mosquitoes Mosquito species

Disease/agent

Distribution

Arboviruses—Togaviridae Tropical Americas Americas, Southeast Asia, eastern Europe United States

Aedes and Culex species Culiseta melanura and Aedes species

Venezuelan equine encephalitis (VEEV) Eastern equine encephalitis (EEEV)

Culex tarsalis and Culiseta species

Western equine encephalitis (WEEV) Arboviruses—Flaviviridae

Aedes aegypti Culex tritaeniorhynchus Culex pipiens, Cx. quinquefasciatus, Cx. tarsalis, Cx. Nigripalpus Culex, Anopheles, and Aedes species Aedes aegypti

Dengue (DENV-1, -2, -3, -4) Japanese encephalitis (JEV) St. Louis encephalitis (SLEV)

Worldwide Tropics Asia, Japan North and South America

West Nile virus (WNV)

Israel, Europe, Russia, northern Africa

Yellow fever (YFV)

South and Central America, Africa

Arboviruses—Bunyaviridae Aedes triseriatus Culex quinquefasciatus

LaCrosse encephalitis (LACV) Rift Valley fever (RVF)

United States Africa

Protozoa Anopheles species An. gambiae, An. funestris, An. stephensi, An. albimanus, An. darlingi, An. darius, An. arabiensis, and others

Malaria Plasmodium falciparum P. vivax P. malariae P. ovale

Worldwide, Tropics and Subtropics

Nematodes Aedes, Anopheles, and Mansonia species Aedes, Culex, and Anopheles species Various mosquito species

Brugian filariasis Brugia malayi Bancroftian filariasis Wuchereria bancrofti Dog heartworm, dirofilariasis, Dirofilaria immitis

to the point at which it can be transmitted to a new host (Eldridge and Edman, 2000; Harwood and James, 1979). Vectorial capacity is the effectiveness of a vector population to transmit a pathogen at a specific time and location. Measuring vectorial capacity means quantifying the key biological interactions of the vector, host, and pathogen under different environmental conditions (Metcalf and Novak, 1994). It generally represents a composite of pathogen virulence, host and vector susceptibility, vector ecology (longevity, feeding preferences, mobility, abundance, diurnal activity, etc.), and the influence of local ecological variables and meteorological parameters (primarily rainfall and temperature) on vector population dynamics (Lehane, 1991; Walker et al., 1996). Understanding vector ecology provides the basis for determining when and where to apply control interventions, and it is particularly important for selective pesticide application (Service, 1993). To determine the spatial and temporal distribution of vector species, effective surveillance techniques are required. Light traps are used for many flying insects,

Southeast Asia Subtropics and Tropics, Worldwide Worldwide

carbon dioxide-baited traps for hematophagous arthropods, chemical- and visual-baited traps for tsetse flies and black flies, oviposition and gravid traps for mosquito eggs and females, dippers for mosquito larvae, and dragging and flagging for ticks (Bidlingmeyer, 1974; Calvin and Gibson, 1992; Reeves, 1990; Service, 1993). Note that most monitoring techniques collect both vector and nonvector species and that accurate taxonomic description of the specimens can be critical. For example, within the genus Anopheles, there are considerable species-specific differences in vector competency and feeding behavior (Curtis and Townson, 1998). Some species avidly blood-feed on humans (anthropophagy); however, their importance in transmission in a particular area may be related to their occurrence inside or outside of human habitats (endophily and exophily, respectively). Refinement of surveillance techniques comes with a better understanding of vector ecology and behavior. Almost all of these techniques provide a relative estimate of population abundance rather than an absolute estimate (number per unit area or volume), and they are best used

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as indicators of the presence or absence of the vector or for seasonal changes in relative abundance (e.g., an increase in oviposition or blood-feeding activity). The transmission cycles of vector-borne diseases can range from simple insect–man–insect transmission (e.g., malaria) to complex interactions with multiple hosts, reservoir species, and vectors that may vary in different ecological habitats (Eldridge and Edman, 2000; Harwood and James, 1979). Vector and host interactions also vary between arthropod developmental stages. For example, ticks generally have larval and nymphal stages associated with small mammal and bird hosts, whereas the adults are often found on larger wild and domestic mammals and humans. Ixodes scapularis, the black-legged tick or deer tick, requires a blood meal for each developmental stage, which takes approximately 2 years from larva to adult (Hinrichsen et al., 2001; U.S. Armed Forces Pest Control, 1990). Each stage feeds only once and each takes several days to ingest the blood. Larvae and nymphs typically become infected with the Lyme disease bacteria, Borrelia burgdorferi, when they feed on infected white-footed mice (Peromyscus leucopus) or chipmunks (Tamias striatus). The adult stage is found on larger animals and is often extremely abundant on white-tailed deer that are usually asymptomatic for the disease. Most Lyme disease cases in humans are associated with the bite of the nymphal stage of I. scapularis, which are difficult to detect because of their small size. Humans primarily contract Lyme disease when they invade the tick habitat either for work or for recreation and due to close proximity of housing to natural areas with an abundance of vectors and small, medium, and large blood hosts. Transmission cycles that occur in the absence of humans, generally under natural or sylvatic conditions, are called enzootic or maintenance cycles, and those involving an increase in transmission to domestic and peridomestic hosts, as well as humans, are called epizootic or epidemic cycles (Eldridge and Edman, 2000). Transmission cycles of a pathogen can vary geographically. St. Louis encephalitis in the United States has at least three distinct vector–pathogen relationships in the western, east central, and southeastern parts of the United States (Monath, 1988; Tsai and Mitchell, 1989). Evidence indicates that genetic variability in the pathogen, some of which may relate to virulence, can be found between large-scale geographic areas, such as the eastern and western United States, as well as smallscale areas, such as differences in SLEV within a single county in Texas (Chandler and Nordoff, 1999; Charrel et al., 1999; Trent et al., 1980). Furthermore, an insect or tick may transmit more than one pathogen, such as I. scapularis, which has been implicated as the main vector for Lyme disease (Borrelia burgdorferi), babesiosis (Babesia microti), and ehrlichiosis (Ehrlichia chafeensis), or Ae. aegypti, which transmits different serotypes of dengue, as well as the yellow fever virus (Calisher and Monath, 1988; Lederberg et al., 1992; Zeidner et al., 2000).

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Epidemiology is the study of the determinants, occurrence, and distribution of diseases in a defined population (Nutter, 1999). It attempts to discover the linkage between the environment and three aspects of disease transmission: host–pathogen, pathogen–vector, and vector–host interactions. The goal is to define the determinants and risk factors in order to develop the most effective measures for prevention and control. For example, control tactics aimed at reducing yellow fever epidemics had only marginal success until it was determined that the mosquito Ae. aegypti transmitted the virus (Karlen, 1995; McGrew, 1995). In general, the more quantifiable information available about the dynamics of the pathogen, vector, and host interactions, the greater the opportunities are for disrupting the transmission cycle (Metcalf and Novak, 1994).

4.3  Impact of arthropods on human health In addition to the emotional burden, vector-borne diseases impose a huge economic encumbrance on families and governments worldwide through lost productivity and health care costs. For example, there are 300–500 million clinical cases of malaria annually, with 1.5–2.7 million deaths each year, including approximately 1 million deaths among children younger than 5 years of age (WHO, 1995; see http://www.who.int/tdr/diseases/malaria/default. htm). In sub-Saharan Africa, malaria accounts for 20% of all childhood deaths. Lymphatic filariasis, another mosquito-borne disease, is second only to mental illness as the world’s leading cause of long-term disability, disfiguring more than 40 million people. Onchocerciasis, or river blindness, transmitted by black flies, places more than 85 million people in Africa, Latin America, and the Arabian Peninsula at risk for visual impairment, blindness, and skin lesions. Sleeping sickness, transmitted by the tsetse fly, threatens 55 million people in 36 countries of subSaharan Africa. In Latin America, up to 18 million people are infected with Chagas disease, a parasitic disease transmitted by blood-sucking true bugs. The chronic stage of Chagas disease can last for years as parasites invade the internal organs. Leishmaniases, caused by flagellate protozoans transmitted by Old and New World phlebotomine sand flies, affects more than 12 million people, damaging internal organs and producing skin lesions and mutilations of the nose and mouth. Approximately two-thirds of the world’s inhabitable land mass is at risk for vector-borne diseases, although it is undeniable that the Tropics have the greatest hardship. There can be little doubt that vector-borne diseases have been instrumental in shaping human history, determining the outcome of wars and limiting human expansion into potentially habitable tropical areas. The reader is referred to several texts about the historical impact of

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vector-borne diseases on human society (Busvine, 1993; Karlen, 1995; Mandell, 1990; McGrew, 1995). It appears that the Eurasian cradle of civilization in the Old World was also the cradle of infectious and vector-borne diseases, which is not surprising considering that the factors favoring domestication of livestock and the development of a stratified society also favored disease transmission (Diamond, 1999). When Europeans started to conquer and colonize the Americas, far more natives died due to the introduction of lethal microbes (smallpox, measles, influenza, mumps, pertussis, tuberculosis, etc.) than on any battlefield. Of approximately 12 infectious diseases of exotic origin that became established in the Americas, 4 were vector-borne – epidemic typhus, malaria, plague, and yellow fever (Diamond, 1999). However, Europeans and their progeny in the New World were not immune to 2 vector-borne diseases introduced through slave trade from tropical Africa, namely yellow fever and malaria (Manson-Bahr and Bell, 1987). Yellow fever was introduced to the New World in the 1500s, and in some cases, cities suffered a 10–15% mortality rate. Colonists and slaves in the 16th and 17th centuries introduced two species of Plasmodium malarial parasites (P. vivax and P. falciparum) into the Americas. During the Lewis and Clark expedition across the Louisiana Purchase in the early 19th century, Peruvian bark powder (containing quinine and quinidine) was considered essential to fight the inescapable fevers of malaria (Ambrose, 1996). Many of these malarious areas had never been visited by Europeans or African slaves; however, the disease had spread through native American trade routes from areas of initial contact. Many of the same vector-borne diseases remain as major impediments to the economic development of tropical areas in both the New World and the Old World. The decline of several vector-borne diseases during the 20th century in the United States was due to a complex mix of variables, including changes in housing conditions, demographics, nutrition, medical diagnosis and treatment, outdoor exposure times, and mosquito abatement. However, North America is far from being free of vector-borne diseases. The vector-borne pathogens in the United States are the arboviral encephalitis viruses transmitted by mosquitoes and the tick-borne borrelial and rickettsial disease agents. Malaria cases do occur in the United States, but they are primarily imported, brought in by travelers from endemic areas. Occasionally, autochthonous transmission does occur, especially in areas near airports and areas where large groups of exposed individuals congregate (Zucker, 1996). Competent vectors of malaria, An. quadrimaculatus and An. freeborni, are still present in the United States. This is true for several arthropod-borne diseases. Outbreaks of dengue in Mexico annually threaten Gulf Coast states such as Texas because they also have the major vector, Ae. aegypti. It is important to realize that although the incidence of major vector-borne diseases has

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been reduced in the temperate zones of Europe and North America, conditions do exist for their emergence and reemergence (Childs et al., 1999; Mellor and Leake, 2000). In the past 40 years, the leading industrialized nations have been overconfident that new technologies, such as vaccinations, pharmaceuticals, synthetic pesticides, and genetic engineering, will eventually manage vector-borne diseases (Gubler, 1998). Although the benefits of technology have not always been as anticipated, they have provided considerable insight into the evolutionary processes governing the co-adaptations of pathogens, vectors, and hosts. Recent events have made us aware that even major industrialized nations can be poorly prepared for the introduction of new pathogens. For example, the West Nile virus (WNV), a flavivirus in the Japanese encephalitis subgroup, was introduced by unknown means into New York City in 1999 and was perceived as two separate epidemics, one in humans and the other in zoo and feral birds [Centers for Disease Control and Prevention (CDC), 1999]. Apparently, by the time the pathogen was correctly identified from the different host sources and vector control was implemented, the epidemic was already declining. Mosquito pools positive for the virus implicated Culex pipiens as the main vector and possibly the overwintering reservoir (CDC, 2000a–c). By the end of the following year, the virus had spread to 12 states, with 21 human cases from Connecticut, New Jersey, and New York; 4323 birds documented to be infected in 12 states; and 60 horse cases in 7 states. The number of mosquito species implicated as potential vectors grew to approximately 12 (CDC, 2001a,b). This human and wildlife threat may severely challenge the public health and vector management infrastructures in the United States.

4.4  Integrated pest management and vector management The concepts and practices of IPM, which were largely developed in response to crop pests, were found to be readily adaptable to arthropod public health pests (Dent, 1995; Kogan, 1998; Metcalf and Novak, 1994). The initial step in vector management is to identify and define, as best as possible, the components of the pest management unit for a specific area. Once the transmission cycle of the pathogen and the life histories of the vector, host, and reservoir or maintenance species are identified, the cornerstone of vector pest management is surveillance. An IPM program can be initiated for almost any public health pest, even with a limited knowledge of the transmission dynamics, by implementing a monitoring strategy. Surveillance determines potential risk, when and where to treat, and the basis for adapting management interventions to a particular area (Service, 1993). Monitoring typically focuses on the incidence of the vector and pathogens. Pathogen surveillance

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may be in vectors, sentinel hosts, or humans (disease surveillance). The detection of pathogens can be broadly classified into direct and indirect methods (Duvallet et al., 1999; Hoeprich et al., 1994; Mandell, 1990; Manson-Bahr and Bell, 1987). Direct pathogen surveillance includes any method that isolates the disease agent, in vivo or in vitro, or some characteristic biochemical or structural component of the pathogen. This includes visual detection, biochemical response to the pathogen, and identification of DNA/RNA sequences or fragmentation patterns. Indirect pathogen surveillance includes methods that rely on an in vivo or in vitro response to the pathogen, including detection of characteristic pathology or antibody response (host serology or various immunoassay methods) (Coyle, 1997). Federal, state, and/or local public health agencies in the United States are usually responsible for pathogen or disease surveillance and measuring trends in disease incidence. However, human disease surveillance is seldom an effective tool for managing an outbreak (Teutsch, 1994). For example, the response to the WNV outbreak in New York City included distribution of repellents to the public, aerial applications of malathion and sumithrin, and a vast public relations effort to warn and advise residents how to avoid pesticide and mosquito exposure. An examination of the 1999 case data and onset dates indicates that the epidemic had already peaked before the majority of these actions were taken (CDC, 1999). In contrast, vector control may have reduced the number of human WNV cases in 2000, although it was unable to contain the spread of the virus (CDC, 2001b). Most mosquito abatement districts (MADs) in the United States focus on monitoring vector species (MADs generally focus on mosquito management; however, their mandate frequently includes other nuisance and vector arthropods and vertebrate pests). The goal of vector management is to implement control techniques to reduce pest abundance below the levels necessary for the transition from enzootic transmission to epizootic or epidemic transmission. Unfortunately, due to the ecological and biological complexities of pathogen transmission, predictive models are few (CDC, 1993; Monath, 1988; Reiter, 1988). Thus, most MADs attempt to prophylactically reduce vector populations without knowing whether they are disrupting a pathogen transmission cycle or not. Typically, they rely on seroconversion of sentinel animal hosts (e.g., chickens for SLEV) or public health bulletins of human cases before they implement emergency control measures, such as ultra-low-volume (ULV) spraying for adult mosquitoes. For several vector-borne diseases, the ability to detect pathogens in low concentration, as well as identify vector and pathogen species and species subgroups by molecular techniques, has revolutionized epidemiological studies and provided vector management groups with an early warning system (Crabtree et al., 1995; DeBrenner-Vossbrinck et al., 1996; Howe et al., 1992; Munstermann and Conn, 1997).

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Monitoring trends in vector populations and disease incidence in humans and animal hosts is the basis for developing and refining predictive epidemiological models, as well as for determining intervention failures and the development of insecticide and/or pharmaceutical resistance. In general, operational failures (e.g., the incorrect amount of pesticide applied to the target vector) are more common than cases of pesticide resistance, especially with the advent of biodegradable carbamates and organophosphorous and pyrethroid insecticides and acaricides (note that not all pesticides within each of these classes can be considered biodegradable) (Bloomquist, 2001; Ware, 1991, 2001; WHO, 1997). However, the possibility of resistance should never be ignored (Brogdon and McAllister, 1998; Hemingway and Ranson, 2000). It has been estimated that 56 of the 66 important Anopheles vectors of human malaria are resistant to three residual insecticides (DDT, lindane, and dieldrin) that were widely used for malaria eradication. At least 31 of these species have additional resistance to the organophosphates malathion and fenitrothion, and another 14 species have multiple resistance to the carbarnate propoxur. Eight Anopheles species are resistant to pyrethroids (Metcalf, 1989a,b). Any management program that relies heavily on a limited number of chemicals for vector abatement should be concerned about resistance (Brown and Pal, 1971). Integrated pest management attempts to minimize the development of insecticide resistance by reducing the selection pressure from a specific chemical agent through the application of physiological, ecological, and/or behavioral specificities (Metcalf, 1998, 1999). Specificity can include the selection of pesticides that are more active against the target organism than nontargets. It also includes rotating chemicals of different classes and/or different vector detoxification mechanisms within and between vector developmental stages. For example, in mosquito control, an early season treatment with an organophosphorous compound may be rotated with a later season treatment with a pyrethroid or by not using the same pesticide for adult and immature stages (e.g., malathion ULV for mosquito adults and temephos liquid formulations for mosquito larvae). Selection pressures can also be reduced by mixing chemical and nonchemical control techniques and by targeting the placement of pesticides to maximize impact on the vector. This includes treating specific areas where there is a concentration of the target vector, such as larval habitats or resting sites of adults, and adjusting the seasonal and diurnal timing of applications to vector population dynamics and flight periodicity. Despite the similarities of vector management to crop pest management, there are significant differences between the two, many of which impact insecticide use. For example, a pathogen transmission cycle may include enzootic and epizootic cycles, involving multiple hosts, reservoirs, and vectors that exhibit considerable habitat, seasonal, and/or bionomic

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variation (Harwood and James, 1979). Transmission cycles may be unknown, not apparent, or difficult to detect and predict. In general, the number of confirmed cases of a vector-borne disease underestimates the number of people infected with a pathogen. Furthermore, action thresholds in vector management (the level of tolerance of disease transmission before a management intervention is taken) are lower than economic thresholds for crop pests (the level of damage tolerance before an intervention is taken). For these reasons, public health management programs rely on long-term and short-term prophylactic treatments to reduce vector populations and eliminate breeding sites before pathogen transmission has been detected, unlike crop pest management (Curtis, 1990; Metcalf and Metcalf, 1993; Mulla, 1994; Reeves, 1990; WHO, 1995). This approach allows the use of natural enemies, source reduction, sanitation and sewage management, vegetation and water-flow (salt, water, and freshwater) management, growth regulators, microbial control agents, and relatively host-specific insecticides (Beidler, 1995; Carlson et al., 1999; Dale et al., 1998; Kramer et al., 1995; Russell, 1999; Wolfe, 1996). Prophylactic interventions are generally habitat and vector specific, whereas emergency interventions tend to rely on insecticides dispensed over broader areas that affect a greater number of nontarget organisms. Insecticides will probably always be an important component of vector management programs because of their ease of application, efficacy, and rapid action. The benefit:cost ratio for pest control is $3 to $5 per $1 invested in agriculture and, for vector control, approximately $2.7 per $1 invested (Metcalf, 1998). Direct treatment of humans with insecticides for vector control is rare, except for ectoparasites such as head lice and scabies mites. Personal protection includes repellents, antibacterials, vaccines (few are available for arthropod-borne diseases), and physical avoidance of the vector. Integrated pest management for public health pests is an areawide problem, involving public and private lands in urban, agricultural, and natural habitats. Therefore, vector abatement programs often require the cooperation of several agencies and/or quasi-legal groups at the local, regional, and national levels. Vector management also deals with several potentially volatile topics, such as human and animal health, pesticide application in urban environments, insecticide impact on feral and domestic wildlife, and modification of human behavior to avoid exposure. Therefore, it typically requires the cooperation of the public and various governmental bodies. In the United States, the MADs are area specific, taxing bodies that focus on monitoring and controlling the vector within a county, suburb, or metropolitan area. Pathogen and/or disease surveillance by local or state public health departments and the CDC assists in detection, trend analysis, standardization of techniques, and training in vector and disease management. Primary caregivers generally control prophylactic and/or therapeutic drugs. The cost of vector management may exceed the

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capabilities of local areas, thus requiring a governmental presence. Unfortunately, those countries at greatest risk for vector-borne diseases are also among the poorest, per capita, countries in the world. The arsenal of insecticides registered for public health pests is relatively limited compared to crop pesticides and, in areas with protected wildlife, consists of less than two or three alternatives. The WHO Pesticide Evaluation Scheme (WHOPES) lists less than 30 insecticides and acaricides for use in public health (http://www.who.int/whopes/quality/en). Despite a WHO evaluation program to identify new pesticides during the past 30 years, less than 5 have been introduced in the past decade. One of the basic concepts of IPM is the sagacious use of insecticides and acaricides. This strategy preserves the pesticide’s efficacy and is essential for maintaining the limited arsenal. An unforeseen dilemma in early attempts at vector eradication was the loss of insecticides and acaricides to resistance. The loss of public health insecticides is also due to a conflict between projected profit, based on anticipated usage, and the cost of meeting federal registration requirements. Legislation that regulates registration of a pesticide, such as the Federal Insecticide, Fungicide, and Rodenticide Act, has produced tremendous benefits for U.S. citizens; however, many chemical industries often conclude that the profit/cost margins are too small (or negative) and do not attempt to register new products or reregister old ones for vector species. Furthermore, the possible elimination of broad classes of insecticides, such as the organophosphorous compounds by the U.S. Environmental Protection Agency (EPA) under the Food Quality Protection Act, has sparked considerable discussion about the future of vector abatement (DiFonzo, 2001). Undoubtedly, the management of arthropod public health pests will become increasingly more important in the future but considerably more challenging.

4.4.1  Noninsecticidal Methods in Vector Management The optimal vector control program has several basic components that include an IPM program planning group, seasonal and full-time employees, facilities and equipment, vector surveillance, disease detection, chemical and nonchemical intervention activities, public education and public relations activities, intergovernmental coordination, data recording and analysis, applied research, emergency contingency plans, and a continuing education component for staff. Most vector management programs employ a wide range of chemical and nonchemical methods to reduce pest arthropod populations before the transmission of a pathogen occurs. The noninsecticidal methods include environmental manipulation; physical, behavioral, and chemical avoidance methods; biological control; the burgeoning field of transgenics and molecular biology; vaccines and therapeutic medicine; and legislative actions. It is beyond the scope of

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this chapter to cover noninsecticidal control methods; therefore, we provide a brief annotated list of possible interventions that vary in compatibility with insecticide treatments: 1. Environmental manipulation or habitat alteration: Environmental manipulation temporarily or permanently changes the environment to eliminate or reduce the number of breeding areas, or the habitat is altered to make the breeding areas less hospitable for the arthropod vector or host (Carlson et al., 1999; Hagmann, 1981; Nayar, 1985; O’Meara, 1992; Shroyer, 1989; TVA, 1947; Wolfe, 1996). Source reduction Removal (drainage) Sanitation and hygiene Temperature treatment of infested materials Source modification Saltwater marsh management and tidal flooding Stormwater and river management 2. Avoidance methods: Several cultural control techniques attempt to physically or chemically prevent or at least reduce the likelihood of contact between humans and the vector (WHO, 1995). Probably one of the most successful cultural controls of flying insects has been the use of physical barriers to make the home a “vectorfree zone” (Gubler, 1998). Quarantines and inspections Timing human activities or place Physical barriers Housing characteristics (screens, air-conditioning, and internal sprays) Bed nets Chemical repellents 3. Biological control: Biological control refers to all of the natural biotic causes of mortality, including predators, parasites, and pathogens (Beidler, 1995; Couch and Bland, 1985; Courtenay et al., 1989; Kerwin and Washino, 1985; Mulla and Chao, 1991; Petersen, 1985; Porter et al., 1993; Rupp and Rupp, 1995; Steelman and Meisch, 1990; Woodring et al., 1996). The terms biological and biorational insecticides are sometimes used to group microbial insecticides with chemical products that are arthropod specific and have minimal environmental impact, such as insect growth regulators (juvenile hormone analogs) or chitinase inhibitors. These are grouped together in the section on chemical control methods. Predators and natural enemies (vertebrates and invertebrates) Parasites (nematodes and arthropods) Pathogens (viruses, bacteria, fungi, and protozoa) Microbial insecticides 4. Transgenics: The refinement of molecular techniques made possible the use of transgenics to control insects (Brousseau et al., 1999; Crampton et al., 1990; James

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et al., 1999; Kokoza et al., 2000; Marshall, 1998; Riebe, 1999). However, some have challenged the utility of transgenics as a method of managing major vector species (Jacobslorena and Lemos, 1995; Spielman, 1994). Carriers of toxic genes Altered vector competence Symbionts 5. Medical treatment: Vaccines are not available for many of the major protozoan and arboviral diseases, although considerable work has been done with malaria vaccine (Brown et al., 1999; Heppner and Ballou, 1998; Kaslow et al., 1999) and experimental vaccines for eastern equine encephalitis (Wilson et al., 1992). Antibiotics successfully treat some protozoan and bacterial pathogens (Strickland, 2000). Antibacterials Vaccines Convalescent care 6. Regulatory or legislative actions: Local, state, and federal laws and policy statements

4.4.2  Chemicals in Vector Management Pesticides are important components of most IPM programs. They are the pharmaceuticals of vector management, and their application should be done on a case-by-case basis with emphasis on doing no harm to nontarget organisms or the environment.

4.4.2.1  Categorizing Insecticides   and Acaricides Pesticides are divided into different classes or categories based on target organism, chemistry, general nature or source, the developmental stage affected, acute and chronic toxicity, general action on pests, biochemical or physiological activity, application method, and formulation. For a more comprehensive review of the mode of action and formulation of pesticides, several texts are recommended (Barlow, 1985; Bloomquist, 2001; CDC, 1981; Curtis, 1990; Goss et al., 1997; Metcalf, 1989b; Metcalf and Novak, 1994; Sukumar et al., 1991; Tomlin, 1994; Ware, 1991, 2001; WHO, 1985, 1988, 1990, 1997, 2001; Worthing and Hance, 1991). The main classification of pesticides is based on the target organism: herbicides, fungicides, rodenticides, bactericides, nematicides, piscicides, avicides, insecticides, and acaricides. Insecticides and acaricides can be grouped by specific target insects, such as termiticides (termites), pediculicides (lice), and miticides (mites). For mosquitoes, the major vector of arboviruses in the United States, insecticides are generally divided into categories based on the developmental stage they affect (i.e., adulticides and larvicides).

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Insecticides and acaricides are commonly grouped into categories based on a combination of their chemistry and general nature or source (Table 4.5). Insecticides may be grouped into inorganic compounds, microbial insecticides, oils and surfactants, compounds of botanical origin, insect growth regulators (IGRs), organochlorines, carbamates, organophosphorous compounds, pyrethroids, pyridine insecticides, and a general group of unclassified insecticides. Each of these groups can be subdivided based on chemistry or mode of action. For example, the organophosphorous compounds are subgrouped based on their specific chemical structure, such as the organophosphates, organothiophosphates, and organophosphonates. Alternatively, the IGRs are subgrouped by the natural compounds they mimic – that is, juvenile hormone analogs and chitin synthesis inhibitors. The U.S. EPA groups pesticides derived from natural materials, such as animals, plants, bacteria, and some inorganics, as biopesticides. Biopesticides have three major categories: transgenic organisms with pesticidal genes, microbial pesticides, and biochemical pesticides such as IGRs and behavior-modifying chemicals. In vector control, the majority of the research has focused on creating a carrier for Bacillus thuringiensis var. israelensis endotoxin genes that provides a slower settling rate, better uptake rate by mosquito larvae, and photostability. Categorizing insecticides by their mode of action yields several different groups, depending on whether the focus is their selectivity, their route of entry, or their site of action. Insecticides may be broad spectrum, impacting many animal groups, or specific, primarily active against the target group of arthropods. Based on route of entry, insecticides are divided into contact, stomach, systemic, or respiratory poisons. They can be persistent (residual insecticides) or short term (biodegradable). Unfortunately, within the biodegradable synthetic insecticides (e.g., carbamates, pyrethroids, and organophosphorus compounds), there are often exceptions to the rule regarding persistence and acute toxicity. Probably the most informative classification system is that based on the biochemical and physiological systems attacked, which provides a better indication of the target specificity. For example, insecticides may be divided into neurotoxins, muscle poisons, metabolic inhibitors, and physical toxicants. These can be even further subdivided. Neurotoxins include chemicals that interfere with nerve transmission (acting on sodium or chloride channels), such as DDT, the pyrethroids, and -aminobutyric acid antagonists; neurotransmitter mimics, such as the nicotinoids and octopamine-related synthetics; and synaptic enzyme inhibitors, such as the carbamate and organophosphorous antiacetylcholinesterases. A detailed discussion of insecticides and their mode of action can be found in Metcalf (1989b) or on the World Wide Web (Bloomquist, 2001; Ware, 2001). Grouping pesticides based on toxicity also yields several categories. A key decision that must be made by the U.S. EPA during the registration process is determining the

Table 4.5  Insecticides by Category for Arthropods of Medical Importancea Inorganic compounds Arsenical insecticides Copper acetoarsenite (Paris green) Fluorine insecticides Barium hexafluorosilicate Cryolite Sodium fluoride Sodium hexafluorosilicate Sulfluramid Microbial insecticides Bacillus thuringiensis israelensis Bacillus sphaericus Lagenidium species Mosquito larvicidal films Larvicidal oils Golden bear oils (GBOs) Bonide Surfactants BVA larvicides Arosurf/Agnique Plant essential oils Botanical insecticides Neem Azadirachtin Pyrethrins Cinerin I, II Jasmolin I, II Pyrethrin I, II Rotenone Ryania Sabadilla Numerous toxic phytochemicals Insect growth regulators Chitin synthesis inhibitors Diflubenzuron Juvenile hormone analogs Fenoxycarb Hydroprene Kinoprene

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Table 4.5  (Continued)

Phoxim

Methoprene

Phosphonate insecticides

Unclassified IGRs Azadirachtin

Trichlorfon Pyrethroid insecticides

Dicyclanil

First generation

Organochlorine insecticides

Allethrin

Chlorinated diphenyls

Second generation

DDT

Bioresmethrin

Methoxychlor

Bioallethrin

Chlorinated benzene

Phenothrin  sumithrin

-HCH, lindane

Resmethrin

Cyclodiene insecticides

Third generation

Chlordane

Fenvalerate

Dieldrin

Permethrin

Endosulfan

Fourth generation

Carbamate insecticides

Bifenthrin

Bendiocarb

-Cyhalothrin

Carbaryl

Cypermethrin

Propoxur

Cyfluthrin

Organophosphorous insecticides

Deltamethrin

Organophosphate insecticides Dichlorvos

Tralomethrin

Naled

Etofenprox

Aliphatic organothiophosphates

Pyridine insecticides Acetamiprid

Malathion

Imidacloprid

Heterocyclic organothiophosphates

Thiacloprid

Coumaphos Pyridine organothiophosphates

Unclassified insecticides

Chlorpyrifos

Borax

Chlorpyrifosmethyl

Crotamiton

Pyrimidine organothiophosphates Diazinon

Diafenthiuron a

Not all chemicals are currently registered.

Pirimiphosmethyl Phenyl organothiophosphates Fenitrothion Fenthion Temephos Fensulfothion Iodofenphos Aliphatic amide organothiophosphates Dimethoate Oxime organothiophosphates

effect a pesticide will have on the environment and nontarget organisms. If a formulated pesticide will not generally cause unreasonable effects, it is classified as “general use.” Conversely, a pesticide that may cause adverse effects on the environment, including injury to the applicator and wildlife, is usually classified as “restricted use.” Restricteduse pesticides may be applied only by or under the direct supervision of a licensed applicator. The level of toxicity of a pesticide is measured by the response to oral, dermal, or respiratory doses of the pure or technical-grade compound. The measurements can be either acute (single dose) or chronic (repeated exposure). Acute toxicity is usually

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reported as the dose of a compound that causes the death of 50% of the test organisms (LD50). Tests on nontarget organisms often include rodents, birds, and fish. Pesticide labels reflect the toxicity of these formulations, based on signal words such as “Danger – Poison” on highly toxic compounds (oral LD50, 50 mg/kg), “Warning” on moderately toxic compounds (oral LD50, 50–500 mg/kg), “Caution” on slightly toxic compounds (oral LD50, 500–5000 mg/kg), and “Caution” on compounds with low toxicity (oral LD50, 5000 mg/kg). Chronic toxicity is the effect of a substance following prolonged and repeated exposure. Dose is expressed as weight of the test substance per unit body weight of test animal (milligram per kilogram) or as weight of the test substance in parts per million in an aqueous solution. For inhalation exposure, dose is expressed as weight of the test substance per unit volume of air (milligrams per liter) or as parts per million per day. For dermal exposure, dose is expressed as weight of the test substance per unit body weight of the test animal or as weight of the substance per unit of surface area (milligrams per square centimeter). In chronic toxicity studies, the no-observed-effect level (NOEL) is the maximum dose used in testing that produces no adverse effects in the test animals (usually mice). The NOEL is usually expressed in terms of the weight of a test substance given daily per unit weight of test animal (milligrams per kilogram per day). The formulated pesticide, the commercial product, seldom contains only technical-grade chemicals. The marketed product typically consists of the active ingredients and sometimes synergists, as well as several inert ingredients that are responsible for improving storage, handling, ease of application, efficacy, and safety. The inert ingredients include surfactants or emulsifiers, stickers, diluents, encapsulants, and other adjuvants. The carrier is the main constituent of the final product and is usually water or oil in the case of sprays, inorganic clays or talcs in the case of dusts or granules, and organic material such as corncob grits in the case of some granules. Most pesticide compounds are relatively insoluble in water and require an emulsifier for them to be mixed with water. Common formulations of insecticides and acaricides include emulsifiable concentrates, wetable powders or water-dispensable powders, flowable suspensions, oil solutions, dusts, granules, aerosols, fumigants, microencapsulated formulations, and baits. Sprays and dusts allow small amounts of pesticide to be applied directly onto the pest or its immediate environment, but both of these formulations tend to have a problem with spray drift. Use of granules is one way to overcome problems with drift if the pesticide does not have to be applied directly onto the pest. Granule formulations are common for several mosquito larvicides, such as methoprene and Bacillus thuringiensis var. israelensis. Insecticides and acaricides for ectoparasites of humans, such as lice and mites, are typically formulated as creams, lotions, or shampoos.

Hayes’ Handbook of Pesticide Toxicology

4.4.2.2  Examples of Pesticides for Public   Health Pests (a) Ectoparasites (Scabies Mite, Head Louse, Body Louse, and Crab Louse) It has been estimated that more than 2 million school-age children develop a lice or scabies infestation each year (Adams, 1996; Brown et al., 1995; Downs, 2000). Scabies is caused by Sarcoptes scabiei mites, which burrow under the skin, leaving a trail of feces and eggs and causing a rash and intense itching. The eggs hatch in 3–8 days, and the larvae move to the surface of the skin, where they molt to two nymphal stages and finally to adults. From egg hatch to adult can take 10–14 days. The scabies mite is generally spread by direct contact and from the clothes or bed linens of an infected individual (Strickland, 2000; van Neste, 1988). Ivermectin is being tested for the treatment of human scabies (Chouela et al., 1999; Yeruham and Hadani, 1998). Lindane is still commonly used to control scabies in Third World countries, along with sulfur compounds and permethrin (Kenawi et al., 1993). The infestation of the body with head or body lice (Pediculus capitis and Pediculus humanus, respectively) is termed pediculiasis, and the condition of having head or body lice is called pediculosis. Infestation with pubic (“crab”) lice (Pthirus pubis) is known as pthiraisis. Head lice and crab lice are not believed to be important vectors, although they may cause intense itching (Burgess, 1990; Harwood and James, 1979; Strickland, 2000). In general, head lice are readily spread by physical contact particularly among schoolchildren, whereas pubic lice are characteristic of adults and spread is often by venereal contact. The body louse, a vector of typhus, trench fever, and epidemic relapsing fever, is less common in the United States than the other two species. A number of insecticides formulated for topical application are available for these pests, including the following: Crotamiton (Eurax) (Burkhart et al., 1998; Ragheb et al., 1995) Lindane (Kwell, Scabene) (Fusia et al., 1987; Robinson and Shepherd, 1980) Malathion (Prioderm) (Brown et al., 1995) Permethrin (Elimite, Nix-OTC) (Fusia et al., 1987; Nassif et al., 1980) RID A-200 Pyrinate (Culver et al., 1988) (b) Acaricides Humans generally encounter ticks on paths and trails in parks or natural areas. The best methods of control are through prevention tactics of the individual. People going to these areas should wear appropriate clothing, including long-sleeved shirts, trousers, and socks. High-top boots should be worn with the trouser legs tucked into either the socks or the boots. Shirts should be tucked into the trousers

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and the long sleeves buttoned. Repellents, such as diethyltoluamide (DEET) or permethrin, applied to the clothing will also provide additional protection from ticks, although permethrin-based products are generally considered better for ticks. Typically, the repellent only needs to be applied as a band around the ankles and waistband. Attached ticks should be carefully removed with tweezers. Pets should be regularly checked for tick attachment and treated as prescribed by a veterinarian. When local premises are infested, a variety of insecticides have been employed successfully to control ticks. The following acaricides have been used for the control of ticks in turf, ornamental, and recreational areas: carbaryl (Sevin), chlorpyrifos (Dursban), cyfluthrin (Tempo), diazinon (Diazinon,), s-fenvalerate (Zema Lawn Spray), fluvalinate (Mavrik Aquaflow, Yardex), and permethrin (Aziz and Osman, 1985; Baxter et al., 1999; Goddard, 1998; Khan, 1999; Monsen et al., 1999; Schuurman, 1988). Several of these pesticides are hazardous to humans and nontarget insects. All pesticides should be applied only as directed on the label. In some cases, licensed personnel are required. A commercially available permethrin product (Damminix) targets the larvae and nymphs of I. scapularis on white-footed mice by filling tubes with insecticide-treated cotton, which mice collect as nesting material (Fehrenbach, 1990; Mejlon et al., 1995). Ticks on the mice are killed, which in turn is supposed to ultimately reduce the number of infected ticks on a treated property; however, mixed results have been reported (Daniels et al., 1991; Stafford, 1992).

4.4.2.3  Insecticides in Mosquito Management Mosquito pesticides are generally classified into two categories: those used to manage the immature stages, or larvicides, and those used to manage the adult stages, or adulticides. An effective pesticide targeting the egg stage, or ovicide, for mosquitoes has not been developed. A variety of larvicides are available to kill mosquitoes, including petroleum and mineral oils, as well as organophosphate compounds, which have been used for decades. In recent years, several effective new insecticides have been used that are less harmful and less persistent in the aquatic environment. These include the microbial insecticides Bacillus thuringiensis var. israelensis (Bti) and Bacillus sphaericus (Bsph) and the IGR methoprene. A new mineral oil coupled with a surfactant is also being used in several areas of the United States. This lightweight surfactant has negligible adverse effects on plants and other aquatic organisms. Table 4.6 lists the current mosquito larvicides registered by the U.S. EPA. There are numerous reports about natural products for mosquito control; however, few have been evaluated for efficacy under field conditions (Dennett et al., 2000; Eckenbach et al., 1999; Lampman et al., 2000; Sukumar et al., 1991).

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It is important that larvicides are applied according to label specifications and only where larvae are present. Control personnel should also be aware that a single larvicide might not be suitable for every mosquito-producing habitat. For example, Bti is only effective against certain species of mosquito, especially floodwater species such as Aedes vexans where the water is relatively clean with little or no organic pollutants. For polluted sites with heavy organic contents, Bsph should be the microbial insecticide of choice. Methoprene is the basic compound for registered IGRs for mosquito control. Because this compound impairs or stops arthropod growth and molting, it should be used only in habitats where mosquitoes are the predominant species or where other beneficial arthropods would not be affected adversely. All insecticides should be used in accordance with the registered label instructions, which include the amount (dose) in pounds or gallons per acre (or kilograms/liters per hectare) and a list of aquatic habitats. A variety of equipment is available for mosquitolarviciding operations, ranging from simple hand-pressurized spray cans to complex equipment mounted on either vehicles or aircraft. The type and quantity of equipment depend on the size of the community and the number and different types of larval habitats. A small community can run an effective program with handheld equipment, whereas larger communities or communities with large tracts of inaccessible habitats such as salt marshes, large flood plains of rivers, or irrigated farmland require more sophisticated application equipment and techniques. More information about larval insecticides and specific products and equipment can be found on the American Mosquito Control Association website at http://www.mosquito.org or from WHOPES at http://www.who.int/whopes. The method of mosquito control most familiar to the public is space spraying, employing vehicle-mounted or aircraft-mounted spray equipment to kill flying adult mosquitoes. A list of registered adult mosquito pesticides is shown in Table 4.6. When space spraying is the principal or only mosquito control activity, good IPM principles are not being followed. In many instances, reliance on space spraying indicates a failure or the lack of proper preventative larvicidal measures against the nuisance or vector mosquito species. Adulticide treatments for mosquitoes must be based on sound ecological and behavioral information about the species to be treated. Mosquitoes exhibit different periods of activity; for example, Ae. vexans and Ae. sollictians are crepuscular with primary activity after sunset and before dawn (Bidlingmeyer, 1974; Horsfall et al., 1973), whereas Ae. triseriatus, Ae. aegypti, and Ae albopictus are active primarily during daylight hours (Hawley, 1988; Novak et al., 1981). The spraying of harborage or aggregation sites can be very effective, as are barrier sprays to minimize mosquitoes moving into an area (Groves et al., 1994; Ham et al., 1999; Mount, 1998; Mount et al., 1996). However, the effectiveness of the treatment is based on good surveillance methods.

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Table 4.6  Insecticides Registered in the United States for Use in Mosquito Control Product

Description of use

Larval control insecticides Temephos

Available in several formulations, including emulsifiable concentrate and on sand or corn cob granules. Higher application rates may be necessary in polluted water. This product is also registered for use against mosquitoes in artificial containers (cans, tires, bird baths, rain gutters, etc.).

Methoprene

Available in liquid, pellet, briquet, and granular formulations. The briquets are suitable for use in small depressions and containers. The insect growth regulator kills mosquito larvae by interfering with the insect hormones that regulate growth. This compound also affects the growth of arthropods and should be used only in habitats as designated on the registered label.

Bacillus thuringiensis var. israelensis (Bti) Bacillus sphaericus (Bsph)

Both Bti and Bsph are microbial larvicides formulated from extracts of bacterial cultures. They are available in liquid, granular, and briquet formulations. Liquid formulations have also been successfully used for ULV ground and air application. Bti is used primarily in unpolluted waters, and Bsph is most effective in polluted and high organic waters

Oils

Larvicidal oils that are formulated to spread and cover a water surface function by preventing the larvae/pupae from reaching the air at the water’s surface or by causing internal toxinosis via the air tube. Particular care should be taken when using an oil in areas where fish, aquatic animals, and plants could be harmed.

Adult control insecticides Naled

An organophosphate insecticide available as a liquid concentrate. It is generally used as a ULV spray from aircraft and ground ULV equipment.

Malathion

An organophosphate available in several liquid formulations. It is generally used as a ULV spray from aircraft and ground ULV equipment.

Pyrethyrins [and piperonyl butoxide, (PBO)]

A botanically derived insecticide generally formulated with PBO. This insecticide has a quick knockdown action on flying mosquitoes.

Permethrin

A synthetic pyrethroid insecticide that can be formulated with a synergist (PBO). It is used primarily as a ULV application or as a perimeter treatment around buildings, parks, etc. to kill resting adults. Pyrethroids have high bee and fish toxicity; follow label directions.

Resmethrin

A synthetic pyrethroid formulated insecticide with PBO. The product has a quick knock-down action on adult mosquitoes. It is more effective at lower temperatures than some other products.

Phenothrin  sumithrin

A synthetic pyrethroid often formulated with other pyrethroids and PBO (used in New York City for West Nile virus vector control).

Adult control treatments must be conducted under environmental and climatological conditions suitable for the application of insecticides using thermal fogs, cold aerosols, or ULV techniques. These treatments should not be conducted if the temperature is below 55°F or above 85°F. These treatments should also be done when wind speed is low (5 mph). When adulticides are sprayed on windy days, turbulence makes a uniform pattern difficult to achieve, resulting in poor control. Also during windy days, the insecticide may be dispersed into areas that are not desirable. Ideally, wind speed should be calculated using an anemometer. The manufacturer of both the adulticiding equipment and the insecticide will provide recommendations for wind, temperature, and, in many cases, relative humidity for determining the limits for ground-applied adulticides.

The ULV technique is the most frequently used method of adult mosquito control [see Mount (1998) for a review of ground ULV applications]. This technique does require special training in the use and especially the maintenance of ULV equipment. Moreover, training for the application of pesticides is mandatory because highly concentrated insecticides are used. For ULV techniques to be effective, the droplets must strike the mosquitoes either in flight or during exposure when resting. Droplets that are too small will not affect active mosquitoes, nor will droplets that are too large because they will settle out of the air too rapidly. Large droplets can also cause spotting on painted surfaces, especially those containing malathion or a corrosive carrier such as oils and synergists. Therefore, the droplets must stay within a specific range in order to achieve maximum

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effect. It is essential that the applicators using this technique consult and adhere to the pesticide’s label and follow the operating instructions for the ULV unit so that the machine is calibrated properly to produce the correct droplet size. Aircraft have been used for many years to apply insecticide dusts, granules, sprays, and aerosols [see Mount et al. (1996) for a review of aerial applications]. Some mosquito control personnel consider it better to use a large volume of low concentrate (e.g., 2 quarts of 55% spray per acre) to obtain better coverage, whereas others believe it is preferable to use a small volume of higher concentrate (e.g., 1 pint of 20% spray per acre). The standard used today assumes that large areas can be covered more economically, especially with a reduction of expensive aircraft downtime, by using small volumes of highly concentrated spray. This led to the development of the ULV spray technique, defined as a method that uses less than 2 quarts of liquid per acre. For example, using malathion at 3 ounces of technical-grade material per acre, 1 gallon (128 ounces) would treat 43 acres. The ULV technique was modified and adapted for the application of 0.5–3 ounces of highly concentrated insecticide per acre for the control of mosquitoes by aerial application. Aerial ULV spraying is used for both nuisance and disease management of mosquitoes. For example, it was used against Culex species during an outbreak of SLEV in Texas in 1966; in New England in 1973, 1974, and 1990 to control mosquitoes transmitting EEEV (Monath, 1988); and for pest management during the major floods in the Midwest in 1993. Aerial applications have also been used very effectively during disasters, for example, in 1989 in South Carolina following Hurricane Hugo and in 1992 in Florida following Hurricane Andrew. For a complete review of aerial applications for adult and larval treatment, see Mount et al. (1996). Because of the highly technical nature of this control method, mosquito control agencies should consult with the manufacturers of the insecticides and application equipment as well as the appropriate regulatory agencies before conducting aerial adulticiding. For more information, see the American Mosquito Control Association website at http://www.mosquito. org or the WHOPES at http://www.who.int/whopes.

4.4.2.4  Personal Protection: Vector and   Pest Repellents Even the most effective mosquito abatement program cannot totally eliminate the nuisances caused by mosquitoes. Therefore, it is necessary at times and in certain environments (picnics, fishing trips, nature trails, natural areas, etc.) to employ personnel protection geared toward minimizing biting mosquitoes. This can be done in a number of ways, including screens in windows and doors, protective clothing, and repellents. Mosquitoes are generally less attracted to white clothes than darker colored clothes and

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clothing that lacks contrasts (i.e., solid colors are better than patterned). Loose-fitting long-sleeve shirts and long pants offer a great deal of protection against mosquitoes and other biting flies. A complete bibliography of repellents for blood-sucking arthropods can be found in Gerberg and Rutledge (2001). Personnel repellents containing DEET are considered to be some of the best mosquito repellents (Elston, 1998; Rutledge et al., 1999). However, their effectiveness depends on the surface area covered and on climatic conditions. On hot, humid summer days, perspiration can effectively wash off the repellent, thus necessitating frequent reapplication. Also, differential repellency by gender has been studied by Golenda et al. (1999). It is not recommended that DEET be applied to the bare skin of preteen children (Garrettson, 1997), and even for adults the applications should be concentrated on clothing and not the bare skin (Chou et al., 1997; Goodyear and Behrens, 1998; Qiu et al., 1998; Young and Evans, 1998). All repellents should be kept away from the eyes, lips, and nasal membranes. Synthetic chemical repellents that are combined with other skin products and sunscreens should be used sparingly. It is very important to read the label and use each product accordingly. For a clinician’s guide to mosquito repellents, see Fradin (1998). Area repellents are available to repulse mosquitoes and other biting flies from a limited area such as patios, gardens, and porches. Many of these products include naphthalene granules, which can be spread on lawn and foliage, and others contain citronella delivered as smoke or from a candle. Many stores carry a variety of electric bug zappers that the manufacturers claim will prevent mosquitoes from biting. These devices also have a “black” (ultraviolet) light that attracts mosquitoes to an electric grid where they are killed. Field trials of these and other devices, such as sonic repellers and mosquito-repellent plants, have been shown not to work against mosquitoes. For more information on these products, see Jensen et al. (2000). Following is a list of synthetic repellents produced commercially or for military, public health emergency applications: diethyltoluamide or N,N-diethyl-3-methylbenzamide (DEET); butopyronoxyl; dibutyl phthalate; dimethyl carbate; dimethyl phthalate; ethyl hexanediol; hexamide; methoquin-butyl; oxamate; and piperidine analogs (Bayrepel). A complete list of phytochemicals and essential oils reported to have repellent activity toward vectors and other arthropods is beyond the scope of this chapter; see Curtis et al. (1990) for a complete review of natural repellents. The following list contains chemicals that have been reported in the literature as repellents, although in some cases their efficacy is unsupported (e.g., the use of vitamin B1). We have made no attempt to rate their effectiveness based on the literature, and they are presented merely as a reference to the variety of natural products claimed to have repellency. In general, natural products are considerably

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less effective than synthetic repellents, such as DEET, at equivalent concentrations. Several chemicals and essential oils do provide relatively short periods of mosquito repellency (Rutledge and Gupta, 1999). Repellency should be considered to have several components. The first component is whether it affects the entire sample population; in laboratory and field tests with mosquitoes, DEET often shows a 100% repellency with initial exposure. A U.S. EPA Scientific Advisory Panel found that “no claim” should be made regarding repellent efficacy for protection against transmission of arthropod-borne disease pathogens. The reasons for this strong response were numerous because most arthropods that interact with humans and animals are capable of transmitting pathogens. Natural products – identified (Hwang et al., 1985; Matsuda et al., 1996; Saxena and Sumithra, 1986): Limonene

Vitamin B1

Linalool

Geraniol

Camphor

Rotundial

Cineole Plant essential oils (Ansari and Razdan, 1995; Barnard, 1999; Curtis et al., 1992; Das et al., 1999; Jensen et al., 2000; Leal and Uchida, 1998; Mulla and Su, 1999; Thorsell et al., 1998): Neem oil

Garlic oil

Aniseed oil

Thyme oil

Geranium oil

Eucalyptus oil

Bergamot oil

Pyrethrum

Lavender oil

Coconut oil

Birchwood tar

Soybean oil

Nutmeg oil

Pine oil

Orange blossom oil Clove oil Cinnamon oil

Pennyroyal oil

Peppermint oil

Citronella oil

For example, we are constantly seeing new worldwide invasions of exotic arthropod-borne diseases such as West Nile virus due to the increasing movement of humans, animals, and domestic goods. Also, individual factors such as proper application, individual variability and susceptibility, and environmental factors all affect the degree of protection afforded by the repellent. In fact, in Gupta and Rutledge (1994), the use of repellents to reduce human– vector contact and reduce the transmission of mosquitoborne diseases was not scientifically proven. Therefore, in

order for the U.S. EPA to rely on the best scientific standards for registration, the use of repellents for reducing arthropod-borne diseases must be determined.

4.4.2.5  Indoor Residual Spraying Indoor residual spraying (IRS) is one of the primary vector control interventions that have been employed to reduce or interrupt the transmission of malaria. However, for many years it has received relatively little attention. Recent data reconfirm the efficacy and effectiveness of IRS in malaria control in countries where it has been implemented. For example, the application of IRS consistently over time in large areas has changed vector distribution and thus altered the epidemiological pattern of malaria in Botswana, Namibia, South Africa, Swaziland, and Zimbabwe. The results of these studies and reports have shown that Anopheles funestus, a major vector, has been either eliminated or reduced to very low levels. The major vector, An. gambiae s.s., which rests and bites mostly indoors, was also controlled primarily due to its proclivity to move into houses. Anopheles arabiensis, which does not move or rest indoors as much as An. gambiae, is less affected by IRS. This was found even in areas where high levels of coverage occurred (Hansford, 1972; Sharp et al., 1990; Southern African Malaria Control, 2000). Anopheles malaria vectors that land and rest inside houses after taking a blood meal (endophilic) are particularly susceptible to control through IRS with contact insecticides. IRS techniques involve coating the walls and other surfaces within a house with a long-lasting residual insecticide. The insecticide of choice should remain active for 1–3 months, killing mosquitoes and other insects on contact. Note that IRS does not directly prevent people from being bitten by mosquitoes because mosquitoes are usually killed after they have fed when they rest on the sprayed surface. The rationale for using IRS is to reduce the vector population to prevent transmission of malaria to other human hosts. Therefore, for this technique to be effective, it must be applied to a large number of households to reduce the vector population to transmission levels or lower. Several ongoing programs have seen dramatic results using IRS as a tactic to reduce malaria transmission in Africa, Asia, and Central America (Arredondo-Jiménez et al., 1993; Charlwood et al., 2001; Doke et al., 2000; Rodríguez et al., 2006; Rowland et al., 2000; Sharp et al., 2007). IRS with DDT and dieldrin was the primary malaria control method used during the Global Malaria Eradication Campaign (1955–1969). Based on its historical impact, several countries have reinitiated DDT as the insecticide of choice for IRS (Gunasekaran et al., 2005; Sharma et al., 2005). Although several African countries have adopted the use of DDT, it is still controversial and has not been endorsed (Casimiro et al., 2007; Kapp, 2004; Sadasivaiah et al., 2007).

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Currently, 12 insecticides are recommended by WHO for IRS, belonging to four chemical groups (1 organochlorine, 6 pyrethroids, 3 organophosphates, and 2 carbamates). The choice of insecticide must be informed by the following considerations: l Insecticide susceptibility and vector behavior l Safety for humans and the environment l Efficacy and cost-effectiveness IRS will only be effective if the target vectors are susceptible to the insecticide in use. The development of resistance to insecticides constitutes a major threat to the chemical control of malaria vectors because it compromises the insecticide’s efficacy. In the past, countries deploying IRS have often been forced to switch to alternative and more expensive insecticides due to the development of vector resistance. Outside Africa, the prevalence and distribution of insecticide resistance in malaria vectors have not been major impediments to insecticide-based interventions, except in some areas of India, the Middle East, and Central America. However, in Africa, the potential threat of resistance to public health insecticides appears to be significant. Resistance to DDT and pyrethroids in major malaria vectors has been found throughout West and Central Africa, in some areas at a high level, as well as in several areas of eastern and southern Africa. Resistance to carbamates has been found in countries of West Africa, with a mechanism that also induces cross-resistance to organophosphates. The selection of resistance in most malaria vectors is thought to be largely the result of past and present use of insecticides in agriculture. The precise operational implications of insecticide resistance are not fully understood. A comprehensive assessment of resistance at the local level must be carried out before planning any IRS program, especially in West and Central Africa. The possibility of insecticide resistance calls for the careful monitoring of the susceptibility of malaria vectors to insecticides throughout the world and the sound management of resistance. There are specific interactions between insecticides and malaria vectors. Some insecticides tend to repel more than to kill vector mosquitoes. Changes in vector behavior induced by insecticides may have important operational implications, and it is important to be aware of them when selecting insecticides for IRS. DDT is the only insecticide that is used exclusively for public health; therefore, unlike with other insecticides, resistance development to it is no longer influenced by other uses such as in agriculture. In the context of resistance management, it is therefore advisable to maintain the use of DDT until a suitable alternative is available. WHO recommends that national governments do the following (see http://malaria.who.int/docs/IRS-position.pdf): 1. Introduce and/or scale up coverage of targeted IRS as a primary malaria control intervention in countries where available data indicate that it can be effective toward achieving malaria targets.

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2. Take all necessary steps to ensure effective implementation of IRS interventions – including selecting the appropriate insecticide, spraying where and when necessary and sustaining a high level of coverage – and to prevent unauthorized or unrecommended use of public health insecticides. 3. Strengthen the managerial capacity of national malaria control programs and improve human, technical, and financial resources for the timely delivery and high coverage of effective interventions including IRS, with adequate monitoring and evaluation. Effective implementation of IRS with DDT or other recommended insecticides should be a central part of national malaria control strategies where this intervention is appropriate. It is implemented with the objective of reducing malaria morbidity and mortality and accelerating progress toward global and national malaria targets. However, there are important considerations that must be taken into account when determining whether to introduce or scale up IRS. In particular, there must be sufficient capacity to deliver the intervention effectively, prevent unauthorized and unrecommended use of public health pesticides, and manage insecticide resistance. Intensified research efforts are needed, for example, to develop new insecticides, long-acting formulations, and improved application technologies. Along with producing IRS manuals and guidelines, WHO will support countries with collecting and analyzing data to determine the potential effectiveness and feasibility of IRS in the national context and with planning and implementing the intervention. WHO requests countries report on coverage and impact as IRS is implemented or scaled up. This position statement is intended for public health policymakers, malaria control program managers, development agencies, development banks, academic and research institutions, and private sector corporations involved in scaling up malaria control programs.

4.4.2.6  Insecticide-Treated Bed Nets Insecticide-treated bed nets (ITNs) are a form of personal protection that has been shown to reduce severe disease and mortality, especially in children, due to malaria. In communitywide trials in several African settings, ITNs have been shown to reduce all-cause mortality by approximately 20% (D’Alessandro et al., 1995; Gimnig et al., 2003; Wiseman et al., 2007). The use of untreated bed nets to form a protective barrier around the people using them has been employed for years to prevent the transmission of malaria via anopheline mosquito bites. However, mosquitoes can feed on people through the nets, and nets that have holes no matter the size provide little protection. The development and implementation of bed nets impregnated with a residual insecticide has greatly enhanced their protective efficacy. The insecticides that have been employed kill the mosquitoes and other insects

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on contact. Some of the insecticides, especially DDT, also have repellent properties that reduce the number of mosquitoes that enter houses (Roberts, 1998). In addition, if high numbers of nets are deployed and used within a community, the numbers and longevity of mosquitoes will be reduced. In these circumstances, even those who do not use a bed net will afford protection from biting anopheline mosquitoes (Etang et al., 2007). However, to achieve such effects, high community coverage is required (Fernando et al., 2008a,b). Early programmatic implementation of ITNs required a reapplication of insecticide at intervals ranging from 6 to 12 months. Moreover, nets had to be retreated every time they were washed. Retreatment was accomplished by simply dipping them in a mixture of water and insecticide and allowing them to dry in an area protected from direct sunlight. This process was considered a major logistical and economic barrier to full implementation of ITNs in countries within malaria (Kroeger et al., 2004; Rafinejad et al., 2008). Currently, several types of nets are available for use in malaria control programs. Nets can vary by size, material, and/or insecticide treatment. The majority of nets are made of polyester, but nets are also available in cotton, polyethylene, or even polypropylene (Sharma et al., 2006; Skovmand et al., 2008). WHO has approved only pyrethroid insecticides for use on ITNs, primarily due to their very low mammalian (human) toxicity but high toxicity to mosquitoes and other insects (N’Guessan et al., 2001; Snow et al., 1999). Pyrethroid insecticides also exhibit a rapid knock-down effect, even at very low doses. In addition, pyrethroids have a high residual effect and do not rapidly break down unless washed or exposed to sunlight, which is minimal because they are used and stored indoors (Dabiré et al., 2006). Several companies have developed long-lasting insecticide-treated nets (LLINs) that retain lethal concentrations of insecticide for at least 3 years. The WHO Pesticide Evaluation Scheme recommends the following LLINs for use in the prevention of malaria: DuraNet (Clarke Mosquito Control; http://duranetmosquitonet.com) Interceptor Net (BASF; http://www.basfpublichealth.com/ products/interceptor.html) NetProtect (Intelligent Insect Control; http://www.insectcontrol.net/netprotect) [also marketed as ICONLife (Syngenta)] Olyset Net (Sumitomo Chemical; http://www.olyset.net) PermaNet (Vestergaard-Frandsen; http://www.vestergaardfrandsen.com/permanet.htm) Detailed reviews by Sexton (1994) and Curtis (1994) provide a comprehensive account of the impact of bed nets on malaria control and as a tactic for vector control. The use of impregnated bed nets by the community has shown good results and is a positive tool for the reduction of malaria (Brieger et al., 1996).

Conclusion The spectacular success of synthetic insecticides in the decade immediately following World War II generated widespread enthusiasm that the major scourges of mankind, such as yellow fever, malaria, and typhus, could be conquered. This in fact did occur, with major reductions in malaria and yellow fever and a significant reduction in other arthropodborne pathogens such as typhus and plague. However, these short-term successes resulted in major increases in disease transmission primarily due to resistance in both the arthropod to insecticides and to drugs by the pathogen. The result of these failures has forced the public health community to reevaluate strategies and has resulted in the development of evidence-based management or integrated disease management. Employing this management strategy, both insecticides or drugs are used in a targeted manner that results in a significant increase in efficacy, environmental safety, and cost-effectiveness. The challenge for the future is to better understand the natural history and bionomics of arthropodborne disease systems and to use this information to target effective control measures. In short, the when, where, and how, coupled with the principle of addressing where the arthropods or pathogens are the most concentrated, immobile, and accessible, should form the basis for future investigations and management programs.

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Chapter | 4  Public Health Pesticides

Organization insecticide irritability test system. J. Am. Mosquito Control Assoc. 15, 60–64. Sadasivaiah, S., Tozan, Y., and Breman, J. G. (2007). Dichlorodiphenyltrichloroethane (DDT) for indoor residual spraying in Africa: How can it be used for malaria control? Am. J. Trop. Med. Hyg. 77, 249–263. Saxena, S. C., and Sumithra, L. (1986). Mosquito repellents from plants. Pesticides 20, 57. Schuurman, H. J. (1988). Diazinon: Acaricide, insecticide with a large spectrum. In “Premieres Journees Veterinaires Africaines, First African Veterinary Days,” June 1987, Hammamet, Tunisia. Office International des Epizooties, Paris. Service, M. W. (1993). “Mosquito Ecology: Field Sampling Methods,” 2nd ed. Elsevier, New York. Sexton, J. D. (1994). Impregnated bed nets for malaria control: Biological success and social responsibility. Am. J. Trop. Med. Hyg. 50(6), 72–81. Sharma, S. K., Upadhyay, A. K., Haque, M. A., Padhan, K., Tyagi, P. K., Ansari, M. A., and Dash, AP. (2006). Wash resistance and bioefficacy of Olyset net – A long-lasting insecticide-treated mosquito net against malaria vectors and nontarget household pests. J. Med. Entomol. 43(5), 884–888. Sharma, S. N., Shukla, R. P., Raghavendra, K., and Subbarao, SK. (2005). Impact of DDT spraying on malaria transmission in Bareilly District, Uttar Pradesh, India. J. Vector Borne Dis. 42(2), 54–60. Sharp, B. L., le Sueur, D., and Becker, P. (1990). Effect of DDT on survival and blood feeding success of Anopheles arabiensis in northern KwaZulu-Natal, South Africa. J. Am. Mosquito Control Assoc. 6, 197–202. Sharp, B. L., Ridl, F. C., Govender, D., Kuklinski, J., and Kleinschmidt, I. (2007). Malaria vector control by indoor residual insecticide spraying on the tropical island of Bioko, Equatorial Guinea. Malar. J. 2(6), 52. Shroyer, D. A. (1986). Aedes albopictus and arboviruses: A concise review of the literature. J. Am. Mosquito Control Assoc. 2, 424–428. Shroyer, D. A. (1989). Salt marsh mosquitoes and disease Bulletin. In “Workshop on Salt Marsh Management and Research,” Vol. 1, pp. 5–6. Florida Anti-Mosquito Association. Skovmand, O., Bonnet, J., Pigeon, O., and Corbel, V. (2008). Median knock-down time as a new method for evaluating insecticide-treated textiles for mosquito control. Malar. J. 7, 114. Snow, R. W., McCabe, E., Mbogo, C. N., Molyneux, C. S., Some, E. S., Mung’ala, V. O., and Nevill, CG. (1999). The effect of delivery mechanisms on the uptake of bed net re-impregnation in Kilifi District, Kenya. Health Policy Plan 14(1), 18–25. Southern African Malaria Control (2000). “SAMC, Malaria Methods: Towards Better Informed Malaria Control in Southern Africa.” SAMC/WHO, Zimbabwe. Spielman, A. (1994). Why entomological antimalaria research should not focus on transgenic mosquitoes. Parasitol. Today 10, 374–376. Stafford, K. C. III (1992). Third-year evaluation of host-targeted permethrin for the control of Ixodes dammini (Acari: Ixodidae) in southeastern Connecticut. J. Med. Entomol. 29, 717–720. Steelman, C. D., and Meisch, M. V. (1990). Biological and genetic control of insect pests of humans, livestock and poultry. Arkansas Farm Res. 39, 3–11. Strickland, G. T. (ed.) (2000). “Hunter’s Tropical Medicines and Emerging Infectious Diseases,” 8th ed. Saunders, Philadelphia. Sukumar, K., Perich, M. J., and Boobar, L. R. (1991). Botanical derivatives in mosquito control: A review. J. Am. Mosquito Control Assoc. 7, 210–237. Teutsch, S. M. (1994). Considerations in planning a surveillance system. In “Principles and Practice of Public Health Surveillance”

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(S. M. Teutsch and R. E. Churchill, eds.), pp. 18–30. Oxford Univ. Press, New York. Thorsell, W., Mikiver, A., Malander, I., and Tunon, H. (1998). Efficacy of plant extracts and oils as mosquito repellents. Phytomedicine 5, 311–323. Tomlin, C. (1994). “The Pesticide Manual: Incorporating the Agrochemicals Handbook,” 10th ed. British Crop Protection Council, Farnham, Surrey, UK. Trent, D. W., Monath, T. P., Browen, G. S., Vorndam, A. V., Cropp, C. B., and Kemp, G. E. (1980). Variations among strains of St. Louis encephalitis virus: Basis for a genetic, pathogenic and epidemiologic classification. Ann. Acad. Sci 354, 219–234. Tsai, T. F., and Mitchell, C. J. (1989). St. Louis encephalitis. In “The Arboviruses: Epidemiology and Ecology” (T. P. Monath, ed.), Vol. 4, pp. 113–144. CRC Press, Boca Raton, FL. TVA (1947). “Malaria Control on Impounded Water,”. U.S. Government Printing Office, Washington, DC. U.S. Armed Forces Pest Control (1990). “Lyme Disease: Vector Surveillance and Control.” Armed Forces Pest Management Board, Walter Reed Army Medical Center, Washington, DC. van Neste, D. J. J. (1988). Scabies. Comp. Pathol. Bull. 20, 3–4. Walker, D. H., Barbour, A. G., Oliver, J. H., Lane, R. S., Dumler, J. S., Dennis, D. T., Persing, D. H., Azad, A. F., and Mcsweegan, E. (1996). Emerging bacterial zoonotic and vector-borne diseases: Ecological and epidemiological factors. J. Am. Med. Assoc. 275, 463–469. Ware, G. W. (1991). “Fundamentals of Pesticides: A Self-Instruction Guide.” Thomson Publications, Fresno, CA. Ware, G. W. (2001). Introduction to insecticides. In “Radcliffe’s IPM World Textbook” (E. B. Radcliffe and W. D. Hutchison, eds.). University of Minnesota, St. Paul. Available at http://ipmworld.umn.edu. Wiseman, V., Scott, A., McElroy, B., Conteh, L., and Stevens, W. (2007). Determinants of bed net use in the Gambia: Implications for malaria control. Am. J. Trop. Med. Hyg. 76(5), 830–836. Wilson, J. H., Gibbs, E. P. J., Calisher, C. E., Buergelt, C. D., Schneider, C. A., Plowright, W., Rossdale, P. D., and Wade, J. F. (1992). Investigation of vaccine-induced tolerance to eastern equine encephalitis in young equids. In “Equine Infectious Diseases VI: Proceedings of the Sixth International Conference,” July 1991, pp. 341–342. R & W Publications, Newmarket, UK. Wolfe, R. J. (1996). Effects of open marsh water management on selected tidal marsh resources – A review. J. Am. Mosquito Control Assoc. 12, 701–712. Woodring, J., Davidson, E. W., Beaty, B. J., and Marquardt, W. C. (1996). Biological control of mosquitoes. In “The Biology of Disease Vectors,” pp. 530–548. University Press of Colorado, Niwot. World Health Organization (WHO) (1972). “Vector Ecology,” Technical Report Series 513. WHO, Geneva. World Health Organization (WHO) (1985). “Specifications for Pesticides Used in Public Health”, 6th ed. WHO, Geneva. World Health Organization (WHO) (1988). “Urban Vector and Pest Control. 11th Report of the WHO Expert Committee on Vector Biology and Control.” WHO, Geneva. World Health Organization (WHO) (1990). “Chemistry and Specifications of Pesticides,” 13th Report of the WHO Expert Committee on Vector Biology and Control, Technical Report Series 798. WHO, Geneva. World Health Organization (WHO) (1992). “Vector Resistance to Pesticides,” 15th Report of the WHO Expert Committee on Vector Biology and Control, Technical Report Series 818. WHO, Geneva. World Health Organization (WHO) (1995). “Vector Control for Malaria and Other Mosquito-Borne Diseases,” Report of a WHO Study Group, Technical Report Series 857. WHO, Geneva.

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World Health Organization (WHO) (1997). “Chemical Methods for the Control of Vectors and Pests of Public Health Importance” (D. C. Chavasse and H. H. Yap, eds.), 5th ed. WHO, Geneva. World Health Organization (WHO) (2001). “Chemistry and Specifications of Pesticides,” Technical Report Series 899. WHO, Geneva. Worthing, C. R., and Hance, R. J. (1991). “Pesticide Manual: A World Compendium,” 9th ed. British Crop Protection Council, Farnham, Surrey, UK. Yeruham, I., and Hadani, A. (1998). Control of human scabies by topical application of ivermectin. Ann. Trop. Med. Parasitol. 92, 627–629.

Young, G. D., and Evans, S. (1998). Safety and efficacy of DEET and permethrin in the prevention of arthropod attack. Military Med. 163, 324–330. Zeidner, N. S., Dolan, M. C., Massung, R., Piesman, J., and Fish, D. (2000). Coinfection with Borrelia burgdorferi and the agent of human granulocytic ehrlichiosis suppresses IL-2 and IFN gamma production and promotes an IL-4 response in C3H/HeJ mice. Parasite Immunol. 22, 581–588. Zucker, J. R. (1996). Changing patterns of autochthonous malaria transmission in the United States: A review of recent outbreaks. Emerging Infect. Dis. 2, 37–43.

Selected world wide web references

http://www.who.int/topics/malaria/en: WHO entry to tropical disease statistics http://www.who.int/tdr/diseases/malaria/default.htm: WHO review of malaria activities http://www-rci.rutgers.edu/~insects/njmos.htm: New Jersey mosquitoes and management fact sheets http://chppm-www.apgea.army.mil: Military pest management handbook http://www.ipmcenters.org/datasources/EPALinks.cfm: EPA, Office of Pesticide Programs http://www.icis.com: Insecticide classification linkinghub.elsevier.com/retrieve/pii/S004835750400118X: Common names of insecticides http://lib.itg.be/tmx.htm: Bibliography of medical entomology texts

http://www.mosquitoes.org/BIORAT.html: An abridged bibliography of selected biorational larvicides for California Mosquito Control, version 5.1 http://www.aphis.usda.gov/vs/nahss/equine/wnv: Animal and Plant Health Inspection Service, U.S. Department of Agriculture, 2000: Summary of West Nile virus in the United States, 1999 http://ace.ace.orst.edu/info/extoxnet/pips/ghindex.html: EXTOXNET, pesticide information profiles http://www.ag.ohio-state.edu/~rich/ag/ag.htm: Mosquito control http://www.cdc.gov/ncidod/dvbid/Arbor/arbofact.htm: Arbovirus fact sheet http://www.mosquito.org: American Mosquito Control Association

Chapter 5

The Changing Role of Insecticides in Structural Pest Control Michael K. Rust University of California, Riverside, California

5.1  Introduction The widespread use and need for pest management in and around structures can be directly attributed to the dislike of insects and arthropods by urban residents. Some insects such as termites and wood-destroying beetles pose a serious economic threat to structures and a few such as cockroaches, fleas, mosquitoes, and ticks may present a real medical or veterinary threat. Even though bedbugs do not vector human diseases, their recent reemergence as a pest worldwide has already dramatically influenced pest management practices and consumer attitudes about structural pest control. Surveys indicate that only 10% of respondents could recognize bedbugs, and individuals 60 years old did somewhat better with 20% identifying them (Reinhardt et al., 2008). Many other urban insects may be classified as occasional intruders and nuisance pests. Surprisingly, some of these such as spiders are considered by homeowners to be a major problem that requires treatment (Rust, unpublished data). In recent years, there has been an increased awareness of the potential hazards of pesticides in urban environments and demand for more integrated pest management (IPM) and so-called “green pest control.” Even though green pest control lacks a clear definition and is largely a marketing tool, it has generated considerable attention. It has revitalized urban IPM, especially strategies that utilize less insecticides, and has stimulated the development of several different training programs for pest management professionals (PMPs). The increased use of containerized and gel baits, insect growth regulators (IGRs), natural products, and alternative strategies such as extreme temperatures, modified atmospheres, and physical barriers is indicative of the public’s increased interest in IPM. This chapter expands on the use patterns of insecticides in and around structures to control the major urban insect pests reviewed by Rust (2001). In recent years, Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

environmental and health concerns have become an important consideration in urban pest management, especially as it relates to water quality issues. The detection of pesticides, particularly pyrethroids, in urban waterways has raised concerns about their use, especially for ant control. Certainly the resurgence of the bedbug, Cimex lectularius L., has focused attention on finding effective residual insecticides and alternative pest control strategies. Wherever possible, information on the efficacy and residual activity of new active ingredients, improved formulations, and alternative strategies will be provided. The chapter concludes with some thoughts regarding the future directions of urban pest management.

5.2  Pest problems: real or perceived The mere presence of insects and arthropods in and around structures is of concern to homeowners and is generally perceived as a pest problem (Rust, 2001). In recent years, there has been a resurgence in the number of bedbugs found worldwide, including Canada (Hwang et al., 2005), Israel (Mumcuoglu, 2008), Italy (Masetti and Bruschi, 2007), Korea (Lee et al., 2008), the United Kingdom (Boase, 2001), and the United States (Anderson and Leffler, 2008). In 2003, Orkin Inc. reported treating 390 cases of bedbugs in 33 different states in the United States (McGinnis, 2004). Public health officials documented 46 locations in Toronto infested with C. lectularius in 2003 and PMPs reported treating another 847 locations (Hwang et al., 2005). Singlefamily dwellings and apartments made up 70 and 18% of the locations treated, respectively. Even in an area endemic for malaria, 33% of those surveyed in Tanzania felt that the tropical bedbug, Cimex hempiterus F., was more troublesome than mosquitoes. Respondents from houses infested with bedbugs always thought them more of a problem than mosquitoes (Temu et al., 1999). In fact, there was a significant 257

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correlation between bedbug infestations and the purchase and use of pyrethroid-treated bed nets. It is easy to see why the topic of bedbugs shares center stage at urban pest management conferences worldwide. In agriculture, the decision to apply an insecticide or treatment strategy (action threshold) is often based on some economic injury level (EIL; Flint and van den Bosch, 1981). In urban settings, economic thresholds have not been established and action thresholds are often based on a perceived economic damage or an aesthetic injury level. The aesthetic injury level (AIL) is highly variable and is based on experience, education, and accompanying attitudes of the target audience (Zungoli and Robinson, 1984). Consequently, the presence of a single bedbug, cockroach, spider, or cat flea indoors may result in an insecticide treatment. However, attempts to establish procedures by which EIL and action thresholds can be established have been proposed (Pinto, 2000; Pinto and Kraft, 2000). Five criteria proposed to establish an EIL are economics, health and safety, esthetics, public opinion, and legal issues (Pinto, 2000). These action thresholds are typically not based on any scientific method and are simply educated guesses. Thus, the action threshold is not zero and the EIL or AIL should be higher (Stejskal, 2002). Preventive treatments should be justified based on the comparison of preventive versus responsive pest control and the cost of forecasted/expected damages (Stejskal, 2003). Justifying preventive treatments and adopting nonzero tolerances will result in fewer calendar-based treatments. If adopted by PMPs, this could have a substantial impact on future urban IPM programs, especially in the reduction of insecticide applications. Pesticides are frequently used and stored in homes (Rust, 2001). This is especially true in low-income housing where cockroach and mice problems commonly exist. In a study of urban minority women, 85% reported pest control measures conducted during their pregnancy and 35% reported their homes being sprayed for pests (Whyatt et al., 2002). Greater than 90% of the pesticides used were to control cockroaches. Most residents in inner city public housing reported using insecticides to control cockroaches and 15% reported using illegal pesticides (Chew et al., 2006).

5.3  Environmental and health concerns In recent years, there has been an increasing concern about pesticides in urban watersheds. In the San Francisco Bay area, it has been suggested that “spraying pesticides on and around buildings to control Argentine ants has historically been among the most problematic pesticide uses for water quality” (TDC Environmental, 2006). The amount of organophosphate insecticides used in urban areas in California significantly decreased from 2000 to 2006, but the use of bifenthrin, cyfluthrin, and permethrin dramatically

Hayes’ Handbook of Pesticide Toxicology

increased (TDC Environmental, 2008). In a suburb of Sacramento, CA, bifenthrin was implicated as the primary cause of toxicity to the amphipod Hyalella azteca Saussure with additional toxicity from cyfluthrin and cypermethrin (Weston et al., 2005). They identified the probable sources of the pyrethroids to structural pest control applications and homeowners’ use of lawn care products. In a study involving 15 creeks in California and 12 creeks around Nashville, TN, 12 of the 15 creeks in California had sediments or water toxic to H. azteca (Amweg et al., 2006). Bifenthrin was suggested as the most likely pesticide causing the toxicity because of its high concentration at each of the sites. Probable sources of the bifenthrin, cypermethrin, and deltamethrin may have been applications made by PMPs. However, eight different products containing bifenthrin were available to homeowners in California and they may have also contributed to its detection in the creeks. Pyrethroids were detected in water runoff and sediments in storm water drains in residential neighborhoods in Sacramento (Weston et al., 2009). Bifenthrin was of greatest toxicological concern with up to 73 ng/l in water and 1211 ng/g in sediment, with cypermethrin and cyfluthrin being of secondary importance. Dry season irrigation was less important than intense storm events in discharging pyrethroids into the storm drains. The total consumer sales of bifenthrin in California in 2005 was 4759 kg AI, 69% of this being purchased from June to September when ants were most active. The total amount applied by PMPs was 19,271 kg AI, 39% being applied from June to September. The patterns of bifenthrin detection were primarily attributed to professional use. One of the major problems with the pyrethroids in these urban systems is that they are primarily particle-bound, smaller particles containing higher concentrations. The concentrations on particles decline slowly over the first year and then persist for a long time (Berger-Preiss et al., 1997). Indoor applications of insecticides are routinely made to control cockroaches. In a study of minority women and pesticides used during their pregnancy, 90% of the pesticides used were for cockroach control (Whyatt et al., 2002). As the level of housing disrepair increased, there was an increase in pesticide use. Air monitoring revealed that diazinon (2– 6010 ng/m3), chlorpyrifos (0.7–193 ng/m3), and propoxur (3.8–1380 ng/m3) were found in residences of all the women monitored in the study. Some women’s exposure to diazinon may have exceeded health-based levels. Urine samples of children in a long-term study were analyzed for the common metabolites of pyrethroids and Lu et al. (2006, 2009) reported continuous exposure to pyrethroids in their diet all year and periodic episodes of high exposure, especially in children in homes where pesticides were applied seasonally. The intensive use of household pyrethroids in the fall and increased consumption of imported produce in the winter and spring may have contributed to the high seasonal levels. Pest control applications increased the levels of pyrethroids such as cyfluthrin, cypermethrin, deltamethrin, and permethrin in house

Chapter | 5  The Changing Role of Insecticides in Structural Pest Control

dust and airborne particles and permethrin and cyfluthrin remained above background levels in house dust for 1 year (Leng et al., 2005). The levels of permethrin and piperyonl butoxide in air samples were highly associated with selfreported use of spray cans and total release aerosols by pregnant women (Williams et al., 2008). Even up to 5 years after chlorpyrifos and diazinon were eliminated for residential use, 92% of the personal air samples contained them. Seventeen different pesticides were analyzed from farm, rural, and urban households (Obendorf et al., 2006). Chloryprifos, resmethrin, and tetramethrin were found in higher levels in urban households. Higher residues were found on carpets compared with smooth surfaces and settled dust. In general, the amount of residues from indoor pest control practices was lower than those in agriculture and horticulture landscapes. In Boston public housing, permethrin, chlorpyrifos, diazinon, and cypermethrin were detected in 100, 100, 98, and 90% of kitchen floor swipes in all units tested, respectively (Julien et al., 2008). Cyfluthrin, restricted to professional use only, was found in 70% of the homes even though they did not hire professional pest control service. Residents were apparently applying the formulated wettable powder to control cockroaches. Appropriately performed pest control applications of pyrethroids such as cyfluthrin, permethrin, cypermethrin, and deltamethrin led to significantly increased pyrethroid metabolite concentrations in occupants at days 1 and 3 post-treatment compared with pretreatment levels (Leng et al., 2003). However, metabolite levels did not exceed published background levels. Studies of German PMPs indicated no expected health effects after occupational applications of pyrethroids as long as label directions were followed (Hardt and Angerer, 2003). Asthma induced by insects and other arthropods is of major importance in structures, especially low-income apartments. The prevalence of cockroach allergens and their mitigation have been reviewed by Arruda (2005), Gore and Schal (2007), and Perzanowski et al. (2008). In a national survey and sampling of residents in apartments in the United States, 11% of living room floors and 13% of kitchen floors had concentrations of cockroach allergen (Bla g 1) exceeding 2.0 U/g (level associated with allergic sensitization) (Cohn et al., 2006). In 3% of living rooms and 10% of kitchens the levels exceeded 8.0 U/g (level associated with asthma morbidity). Elevated concentrations were associated with high-rise apartments, homes built before 1940, urban areas, low-income households, and multifamily structures. Residents (86%) reporting cockroach problems had considerably higher levels of allergen, 58% having 2.0 U/g and 38% having 8.0 U/g in their homes, than did those not reporting cockroaches. In low-income housing, 81% of the apartments were infested with cockroaches, mice, and other insects (Wang et al., 2008). Kitchen dust samples in 98% of the

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apartments sampled had detectable levels of cockroach allergen (0.4 U/g Bla g 1), 52% had 2 U/g, and 33% had 8 U/g. Of the 1173 residents in those units, 13 and 9% had medically diagnosed asthma and allergy, respectively. In a study of New York City public housing, cockroaches were found in 77% of the apartments and 15% of the residences had used illegal pesticides, cockroach chalk, and Tempo (cyfluthrin registered for professional use only; Chew et al., 2006). Dust samples containing cockroach allergen (Bla g 2) were 8 U/g in 36% of the kitchens and 15% of the bedrooms sampled. Apartments with one or more asthmatic patients were characterized by beds with high cockroach allergen, cockroaches in the kitchen, and resident reports of seeing cockroaches. Most residents reported using cockroach baits, but 66% indicated that they had used total release aerosols or foggers, insecticidal chalk, sprays, and Tempo. Environmental interventions including sealing cracks and small holes, applications of gel baits to cracks and crevices, boric acid, and vacuuming cockroach fecal material and debris in a Boston public housing project resulted in clear reductions of cockroach allergens (Levy et al., 2006). The greatest reduction in cockroach allergens corresponded to the greatest improvement in asthma health conditions in the apartments. The Bla g 1 allergens were reduced by 71 and 53% in the kitchen and bedroom, respectively, and Bla g 2 allergens by 86 and 70%, respectively, within 6 months (Peters et al., 2007). After 6 months, the concentrations of allergens began to increase indicating that the intervention must be sustained. In a study comparing a baiting program conducted by university-based entomologists and service provided by PMPs, the baiting program provided 90% reductions in cockroaches trapped, whereas the PMPs achieved between 62 and 81% reductions (Sever et al., 2007). There were significant reductions in allergens (Bla g 1) at all locations in homes baited by university personnel. Even though there was an 83% reduction in cockroaches in kitchens in PMP treatments, the level of allergen only decreased by 35.7%. Clearly, very substantial reductions or near elimination of cockroaches is needed to make significant reductions in the allergens present. In addition to living with cockroaches in housing, occupational risks also occur. In a survey of seamen, 68.3% reported exposure to cockroaches on board ship (Oldenburg et al., 2008). Of them, 26.9% were cockroach sensitive to a skin prick test. Seamen from tropical countries (37.3%) were significantly more sensitive than seamen from temperate countries (21.3%). A new source of indoor allergens in structures is the invasive Asian ladybird beetle, Harmonia axyridis Pallas (Nakazawa et al., 2007). Adult ladybird beetles enter structures in the fall, frequently overwintering in large numbers. Patients primarily exposed to allergens from the German cockroach, Blattella germanica (L.), had a significant cross-reaction to two proteins isolated from H. axyridis.

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5.4  Insecticide applications 5.4.1  Nonresidual Insecticides When insecticides are applied to kill existing infestations and do not provide long-lasting effects, they are typically classified as nonresidual applications (Anonymous, 1989). These can be defined as follows: Nonresidual insecticides are those products applied to obtain insecticidal effects only during the time of treatment and are applied either as space treatments or as contact treatments: a. Space treatment is the dispersal of insecticides into the air by foggers, misters, aerosol devices, or vapor dispensers for control of flying insects and exposed crawling insects. b. Contact treatment is the application of a wet spray for immediate insecticidal effect. The nonresidual insecticides are most typically applied as aerosol or spray formulations of allethrin, esfenvalerate, resemethrin, and synergized pyrethrins (Braness, 1997). Considering the popularity of aerosol sprays, there is a paucity of data concerning their efficacy. A review of some earlier studies is provided by Rust (2001). Alternative low-impact chemical treatments are popular especially among those interested in so-called green pest control. High-viscosity foams produced with air or CO2 killed significantly more German cockroaches than did lowviscosity foams (Choi et al., 1997). Baldwin and Koehler (2007) reported that contact sprays of various commercially available dishwashing liquids and household cleaners were toxic to adult male German cockroaches. Topical sprays of a 0.54% solution provided 50% kill of adult males. In recent years, the use of the exempt products that contain ingredients listed in Table 5.1 has become popular with PMPs. In contact toxicity studies against German cockroaches, pulegone, camphor, and verbenone were comparable to permethrin (Jang et al., 2005). Eucalyptus, marjoram, pennyroyal, and rosemary oils had more pronounced contact and vapor activity against human head lice, Pediculus humanus capitis De Geer, than did -phenothrin and pyrethrum (Yang et al., 2004). In vapor phase toxicity tests, verbenone, -thujone, thymol, -terpineol, camphor, linalool, and marjoram oil were toxic to cockroaches, but considerably less active than dichlorvos. Monoterpenoids such as 1,8-cineole, anisole, limonene, and -pinene showed vapor activity against eggs and adult P. humanus capitis (Toloza et al., 2008). Deposits of basil, citronella, lemon, peppermint, and tea tree oils were not highly toxic to Argentine ants, Linepithema humile (Mayr), and red imported fire ants, Solenopsis invicta Buren, but deterred them from crossing barriers (Wiltz et al., 2007). Only citronella oil provided 50% kill of Argentine ants in 34 min and 100% kill within 24 h. The mode of action of the essential oils remains unclear. Enan (2001) reported that eugenol, -terpineol, and

Table 5.1  Active Ingredients Which May be in Minimum-Risk Pesticide Products Exempted under Section 25(b) of FIFRA Castor oil (U.S.P. or equivalent)

Linseed oil

Cedar oil

Malic acida

Cinnamona and cinnamon oila

Minta and mint oila

Citric acida

Pepperminta and peppermint oila

Citronella and citronella oil

2-Phenethyl proprionate (2-phenylethyl proprionate)

Clovesa and clove oila

Potassium sorbate

a

Corn gluten meal

Putrescent whole egg solids

a

Corn oil

Rosemarya and rosemary oila

Cottonseed oila

Sesamea and sesame oila

Dried blood

Sodium chloridea

Eugenol

Sodium lauryl sulfate

a

Garlic and garlic oil

a

Soybean oil

Geraniol

Thymea and thyme oila

Geranium oil

White peppera

Lauryl sulfate

Zinc metal strips a

Lemon grass oil a

These active ingredients are exempt for use on all food commodities from the requirement of a tolerance on all raw agricultural commodities at 40 CFR 180.1164(d) (EPA, 2009).

cinnamic alcohol were neuro-insecticides and their toxicity was species-dependent. The octopaminergic system mediated the insecticidal activity of eugenol and -terpineol, but not cinnamic alcohol. Eugenol has been reported to affect octopamine receptors and mimicks octopamine, thereby increasing cellular calcium levels, and causes toxicity (Enan, 2001). However, Price and Berry (2006) reported that eugenol and octopamine had opposing effects on dorsal median neurons in the American cockroach, Periplaneta americana L., inferring that eugenol must interact on a different subtype of octapmine. Geraniol and citral showed some similarities to octopamine. Inhibition of acetylcholinesterase (AchE) has been proposed as a mode of action for the essential oils. However, Picollo et al. (2008) found the vapor activity of 1,8-cineole and toxicity to P. humanus capitis greater than dichlorvos but found no correlation with inhibition of AchE activity. Even though there has been considerable interest in these essential oils and compounds exempt from U.S. Environmental Protection Agency (EPA) registration, there has been very little published about their use in pest management programs and effectiveness. With limited residual

Chapter | 5  The Changing Role of Insecticides in Structural Pest Control

activity, additional structural modifications and cultural controls may be necessary to provide satisfactory control, especially for PMPs providing every other month or quarterly pest control.

5.4.2  Residual Insecticides Insecticides applied to provide control over a longer period of time are typically referred to as residual insecticides (Anonymous, 1989). These are defined as follows: Residual insecticides are those products applied to obtain insecticidal effects lasting several hours or longer and are applied as general, spot, or crack and crevice treatments: a. General treatment is an application to broad expanses of surfaces such as walls, floors, ceilings, or as an outside treatment. b. Spot treatment is an application to limited areas on which insects are likely to occur but will not contact food or utensils or ordinarily contact workers. These areas may occur on floors, walls, and bases or undersides of equipment. For this purpose a “spot” will not exceed 2 square feet. c. Crack and crevice treatment is an application of small amounts of insecticides into cracks and crevices in which insects hide or through which they may enter the building. Such openings commonly occur at expansion joints, between different elements of construction, and openings that may lead to voids such as hollow walls, equipment legs and bases, conduits, motor housings, and junction or switch boxes. Residual insecticides have been the primary chemicals used by PMPs because they allow for less frequent visits and applications. Consequently, they are routinely applied according to a schedule in a preventive manner. Typically insecticides only account for 3–5% of the costs of treatments provided by PMPs and thus labor and travel are much more important in reducing costs.

5.4.2.1  General Treatments Sprays are frequently applied around structures as barriers to prevent the accumulation of insect pests and their movement into structures. Typically, treatments consist of high volumes (0.5–1.0 gal/100 ft2) of low concentrations of insecticides applied in bands 2.4–3 m wide on and adjacent to the structure (Klotz et al., 2008; Rust, 2001). In the past few years, this practice has been questioned because of the potential for insecticide runoff from hard surfaces into urban surface waters. Consequently, more targeted and site-specific application techniques for ant control are being promoted (Klotz et al., 2007, 2009). Physiological insecticide resistance is still an important problem in situations in which residual insecticides are widely applied for the control of German cockroaches. Strains of B. germanica collected in Cuba were resistant to

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malathion, deltamethrin, and cypermethrin (Pantoja et al., 2000). A field-collected strain of German cockroach was highly resistant to permethrin and deltamethrin, had low cross-resistance to imidacloprid, and had no cross-resistance to fipronil (Wei et al., 2001). German cockroaches collected from 29 different locations in North Carolina exhibited resistance to cyclodienes and cross-resistance to fipronil, and there was a direct relationship between survivors exposed to dieldrin and fipronil (Holbrook et al., 2003). These cockroaches were resistant to ingested fipronil bait. In urban areas of Taiwan, patterns of resistance in B. germanica populations were propoxur  chlorpyrifos  cypermethrin in hospitals, and propoxur  cypermethrin  chlorpyrifos in houses (Pai et al., 2005). Resistance patterns reflected the overall insecticide use in both hospitals and homes. Two field-collected strains in Denmark were resistant to dieldrin and the most resistant strains were homozygous for A302S mutation, imparting 1000-fold resistance to dieldrin and 15-fold resistance to fipronil (Kristensen et al., 2005). In Iran, 11 field-collected strains of German cockroaches were resistant to permethrin and showed no cross-resistance to fipronil (Nasirian et al., 2006). Field-collected strains of B. germanica were shown to have multiple resistance mechanisms to permethrin, namely kdr-type and metabolic resistance (Limoee et al., 2007). The presence of pyrethroid resistance must be taken into account when developing an IPM program to control German cockroaches. The use of inorganic dust formulations, silica aerogels, and boric acid to treat wall voids, attics, and other areas likely to harbor cockroaches and silverfish has been advocated for decades (Ebeling, 1995). Our understanding of the mode of action of boron remains unclear, and a review of the literature on boron toxicity in insects is provided by Gentz and Grace (2006) with special emphasis on termites. Subterranean termites were confined to borate-treated wood surfaces for 5 days, then removed, and placed on untreated wood. They excreted or metabolized about 60% of the boron and survived the exposure (Gentz and Grace, 2008). This clearly supports the use of borate-treated lumber in a preventive role in IPM programs but not as a remedial treatment. Even though diatomaceous earth and silica aerogel are principally silica, the high sorptivity of insect wax by silica aerogels makes them much more effective as a desiccant than diatomaceous earth. Diatomaceous earth principally acts as an abrasive (Ebeling, 1995). Diatomaceous earth treated with dichlordimethyl-saline significantly increased the insecticidal activity against German cockroaches even at high humidities (Faulde et al., 2006). Contrary to common knowledge, insecticidal dust deposits including boric acid, Drione, silica gel, silica gel with pyrethrins, and organic dusts can be used in high-humidity environments (Appel et al., 2004). The application of perimeters sprays, especially for the control of ants, is a widespread and standard

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approach to controlling pests outdoors. Barrier sprays of cyfluthrin, imidacloprid, and fipronil gave varying degrees of control against several species of ants (Scharf et al., 2004). However, this approach has become a major concern because of potential pesticide runoff into urban waterways. Reduced application techniques to ant trails and structural guidelines provided equivalent control to full perimeter treatments and reduced the volume of spray applied by 67% (Klotz et al., 2007, 2009). The effectiveness of perimeter sprays against ants is enhanced by delayed toxicity and horizontal transfer of insecticides by necrophoresis (Soeprono and Rust, 2004a). When ants exposed to fipronil die and are removed by nestmates, lethal doses of insecticide are transferred to nestmates (Choe and Rust, 2008; Soeprono and Rust, 2004b). This horizontal transfer clearly explains the outstanding results achieved with perimeter sprays of fipronil (Klotz et al., 2007, 2009). Perimeter spray applications around structures similar to those applied against ants with microencapsulated cyfluthrin resulted in low but measurable levels of cyfluthrin up to 9.1 m from the structure (Stout and Leidy, 2000). No cyfluthrin was measured in the air indoors, but it was detected on some indoor surfaces. Treated soil around the structure was probably the source of the indoor residues. Application techniques that reduce the likelihood that pesticides move from the target site will become the new standard in the future.

5.4.2.2  Spot Treatments More precise applications of insecticides are often necessary in environmentally sensitive areas such as food preparation areas, hospitals, offices, nursing homes, computer facilities, and museums. The area to be treated and the amount of residual insecticide to be applied are often important considerations. In areas where food is being prepared, pest management strategies require the judicious and safe use of insecticides. The following definition specifically delineates such areas (Anonymous, 1989): A food handling establishment is an area or place other than a private residence in which food is held, processed, prepared. and/or served. a. Nonfood areas of food handling establishments include garbage rooms, lavatories, floor drains (to sewers), entries and vestibules, offices, locker rooms, machine rooms, boiler rooms, garages, mop closets, and storage (after canning or bottling). b. Food areas of food handling establishments include areas for receiving, serving, storage (dry, cold, frozen, raw), packaging (canning, wrapping, boxing), preparing (cleaning, slicing, cooking, grinding), edible waste storage, and enclosed processing systems (mills, dairies, edible oils, syrups).

Hayes’ Handbook of Pesticide Toxicology

Even though the treatment of these sensitive areas represents a significant income to PMPs, there has been very little published on the efficacy of spot treatments. Nothing better highlights this situation than the lack of data regarding spot treatments of insecticides for the control of bedbugs. The resurgence of C. lectularius has stimulated a renewed interest in finding residual insecticide treatments that can be applied inside structures. Continuous exposure on residual deposits of a probable susceptible strain of bedbugs provided the following lethal times required to kill 50%: -cyhalothrin  bifenthrin  deltamethrin  permethrin  chlorfenapyr (Moore and Miller, 2006). None of the insecticides was repellent. By contrast, Romero et al. (2009) found that bedbugs avoided filter paper treated with deltamethrin. However, a field-collected strain of C. lectularius resistant to deltamethrin did not avoid harborages treated with deltamethrin. Chlorfenapyr deposits failed to prevent bedbugs from mating and laying eggs (Moore and Miller, 2006). Barriers of deltamethrin or chlorfenapyr failed to prevent bedbugs from reaching feeding sources (Romero et al., 2009). The development of insecticidial resistance and worldwide human travel and commerce may be the primary factors leading to the resurgence of bedbugs. Resistance to permethrin and -cypermethrin has been documented in the tropical bedbug, C. hemipterus, possibly resulting from bed net programs to control malaria (Myamba et al., 2002). Antimalarial programs in Sri Lanka may have been an important factor contributing to broadscale resistance in C. hemipterus (Karunaratne et al., 2007). Karunaratne et al. (2007) reported populations resistant to DDT, malathion, permethrin, and deltamethrin. A field-collected strain of C. lectularius tested against deltamethrin required 343 h to provide 50% kill compared with only 1 h in the susceptible strain (Moore and Miller, 2006). Considerably more research is needed to find effective chemical alternatives to the pyrethroids. Various microencapsulated formulations (CS) were tested against B. germanica for their bioavailability from porous and nonporous surfaces (Stejskal et al., 2007). Larger microcapsules were more readily available on porous substrates than were smaller microcapsules. The research suggests that increasing the size of the microcapsules would improve the toxicity of those pyrethroid CS formulations applied to porous substrates. This would be especially important for outdoor barrier treatments and indoor treatments on greasy and oily surfaces (Rust, 1995).

5.4.2.3  Crack and Crevice Treatments It is generally accepted that crack and crevice applications reduce the amount of insecticides in the indoor environment compared with aerosols and sprays. Crack and crevice applications of chlorpyrifos in the kitchen resulted in transport of chlorpyrifos to all locations in the house because it is a semivolatile compound with a vapor pressure of 1.7  105  mm  Hg at 25°C (Stout and Mason 2003).

Chapter | 5  The Changing Role of Insecticides in Structural Pest Control

Air conditioning in the structure may have contributed to the movement. The pin-stream application also produced localized splashing and higher deposition of chlorpyrifos. As expected, a total release application of a similar amount of chlorpyrifos produced much higher amounts throughout the structure. Pyrethroids are not very volatile and translocation may be more influenced by movement of particle-bound residues. When applied indoors according to label directions, concentrations of permethrin and deltamethrin were low in the air and mainly particle bound (Berger-Preiss et al., 1997). Smaller particles contained higher concentrations of each pyrethroid. The concentrations on particles declined slowly over the first year and then persisted for a long time. In a year-long study comparing a conventional monthly baseboard spray (0.025% cyfluthrin) and crack and crevice treatment (boric acid aerosol) with an IPM approach incorporating vacuuming and monthly or quarterly applications of baits (2.15% hydramethylnon bait) and IGR devices (90.6% hydroprene), Miller and Meek (2004) found that the conventional treatment cost $1.05 per unit treated whereas the IPM approach was $4.06. However, the average trap catches (15.3) in the conventional routes were significantly higher than the IPM route (9.2) at 12 months. In another study comparing spray and crack and crevice treatments with baiting programs,

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monitoring with traps clearly revealed that most of the conventional spray treatments were totally unnecessary because of extremely low numbers of cockroaches trapped in the schools (Williams et al., 2005). Overall the costs of the services were similar and the IPM eliminated using propetamphos sprays.

5.5  Soil treatments for subterranean termites For nearly 50 years, the application of organochlorine insecticides such as chlordane, heptachlor, aldrin, and dieldrin to soil was the primary preconstruction treatment or remedial strategy to protect structures from subterranean termites. These insecticides provided an inexpensive and persistent barrier to termites (Su and Scheffrahn, 1998). With the loss of organochlorine chemistry, there has been a progression of insecticides registered for termite control as partially reviewed by Rust (2001) and Kard (2003). Over the past 20 years, the water solubility and contact toxicity of many of these insecticides against subterranean termites have increased, whereas the Kow (octanol water coefficient), Koc (soil organic carbon-water coefficient), and soil half-lives have decreased (Table 5.2). Consequently, the paradigm about soil applications

Table 5.2  Insecticides that have been Registered for Preventive and Remedial Applications to Soil for the Control of Subterranean Termites Insecticide Chlordaneb c

Chlorpyrifos d

Isofenphos

e

Permethrin

Classa

Water solubility (g/l)

Log Koc

Log Kow

Soil half-life

OC

0.0001 

4.19–4.39

2.78

~4 years

OP

0.002 

3.78

4.7–5.11

11–141 days

OP

0.018–0.024 

2.89

3.30

59–127 days

PY

0.0002 

4.93

2.88

21–42 days

f

Cypermethrin

PY

0.000009 

5.20

6.6

8–16 days

g

PY

0.00002 

3.72

6.22

15–90 days

Esfenvalerate h

Bifenthrin

PY

0.0001 

5.10–5.48

6.0

65–156 days

i

NE

0.51 

2.12–2.49

3.7

48–190 days

j

CP

0.00014 

3.32–3.69

4.83

146–730 days

PP

0.0019 

2.92

5.0

122–128 days

NE

3.5–4.2 

2.12–2.43

0.8

1–8 days

Imidacloprid

Chlorfenapyr k

Fipronil

l

Acetamiprid a

OC  organochlorine; OP  organophosphate; PY  pyrethroid; NE  neonicotinoid; CP  cyanopyrrole; PP  phenylpyrazole. extoxnet.orst.edu/pips/chlodan.htm; www.eps.gov/ogwdw/dwh/t-soc/chlordan.html c pmep.cce.cornell.edu/profiles/extoxnet/carbaryl-dicrotophos/chlorpyrifo-ext.html d www.ars.usda.gov/util/download.cfm?file  SP2userFiles/ad_hoc/12755100DatabaseFiles//isofenphos.txt e pmep.cce.cornell.edu/profiles/extoxnet/metiram-propoxur/permethrin-ext.html f pmep.cce.cornell.edu/profiles/extoxnet/carbaryl-dicrotophos/cypermet-ext.html; www.flouridealert.org/pesticides/msd/flutriafol.armour.c.seed.dress.pdf g pmep.cce.cornell.edu/profiles/extoxnet/dienochlor-glyphosate/esfenvalerate-ext.html h www.pw.ucr.edu/textfiles/bifentn.pdf; pmep.cce.cornell.edu/profiles/extoxnet/24d-captan/bifenthrin-ext.html i pmep.cce.cornell.edu/profiles/extoxnet/haloxyfop-methylparathion/imidacloprid-ext.html; www.cdpr.ca.gov/docs/emon/pubs/fatememo/imid.pdf j pmep.cce.cornell.edu/profiles/insect-mite/cadusafos-cyromazine/chlorfenapyr/chlorfenapyr_DECletter_102.html k npic.orst.edu/factsheets/fipronil.pdf l www.epa.gov/opprd001/factsheets/acetamiprid.pdf

b

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of insecticides to protect and remediate termite infestations has dramatically changed. Remedial treatments provide shorter residual control, but their activity against termites has increased. For decades the U.S. Forest Service has been testing insecticides in two different field tests as potential soil treatments against subterranean termites (Kard, 2003; Wagner et al., 2003). Seven different organophosphate and pyrethroid termiticides were applied to two different soil types in Mississippi and soils were extracted and analyzed over 5 years (Jarratt et al., 2004). The initial concentrations at both sites were two times higher than expected according to label directions. Only isofenphos failed to meet minimum thresholds at 5 years. However, with the advent of the new slow-acting nonrepellent insecticides, the utility of these two study designs has come under closer scrutiny. In other field situations, the residual activity of some of the insecticides has been considerably less than reported in the U.S. Forest Service studies. Organic matter content, silt and clay proportions, pH, and cation exchange capacity affected the bioavailability of imidacloprid to subterranean termites (Ramakrishnan et al., 2000). It was not detectable at 0.05 ppm at several homes 1 year after treatment with 0.05% imidacloprid (Osbrink et al., 2005). Soil residual studies in Arizona found that insecticides degraded as follows: chlorpyrifos  imidacloprid   bifenthrin  permethrin  zeta-cypermethrin (Baker and Bellamy, 2006). The registered termiticides provided between 2.85 and 5.4 years’ activity. In field studies in Australia, soils treated with bifenthrin and chlorfenapyr were the most persistent, losing  50% over 12 months (Horwood, 2007). Chlorpyrifos, fipronil, and imidacloprid were least persistent, losing 99, 67–96, and 50–95%, respectively. In simulated field studies, imidacloprid levels declined from 84.5 to 7.5 ppm in 6 months and the insecticidal activity of deposits down to 15 cm decreased dramatically after 9 months (Peterson, 2007). Preconstruction treatment rates of bifenthrin, chlorpyrifos, and imidacloprid applied at two locations in Mississippi were tested for longevity and activity (Mulrooney et al., 2006). Bifenthrin and chlorpyrifos maintained their insecticidal activity for at least 48 months, whereas the imidacloprid significantly declined within 12 months. Their study showed that most of the insecticide and contact with termites occurred within the top 2.5 cm of the surface. Richman et al. (2006) found that elevating soil pH with Portland cement affected residual activity of insecticides as follows: imidacloprid  fipronil  chlorpyrifos  bifenthrin  permethrin  cypermethrin. As soil organic matter increased, the bioavailability of fipronil decreased (Mulrooney and Gerard, 2007). Saran and Kamble (2008) found that the half-lives of bifenthrin, fipronil, and imidacloprid increased as the concentration applied to the soil increased and there was an inverse relationship between concentration applied and the time required to kill 50 and 90% of the termites. The

Hayes’ Handbook of Pesticide Toxicology

authors suggest that as the concentration increases the adsorption coefficient decreases, increasing the amount of toxicant available to kill termites (Kamble and Saran, 2005). In general, the residual activity of the newer nonrepellent termiticides is significantly less than previously registered organochlorine, organophosphate, and pyrethroid insecticides. With the advent of the slow-acting nonrepellent chemistries, it was suggested that the insecticides were transferred horizontally through the colony providing area-wide effects and colony kill (Potter and Hillery, 2001). Ninetythree percent of monitoring stations 0.3–4 m away from fipronil-treated structures had no termite activity or had dead termites. It was suggested that the use of nonrepellent slow-acting termiticides might make dogmatic mandatory full treatments of interior perimeter expansion joints and other indoor sites unnecessary (Potter and Hillery, 2002). There has been very little direct evidence to support this claim. The perimeters of buildings were treated with 0.05% imidacloprid, and termites collected in traps within 1–3 m from the treated zone were not affected by the treatment (Osbrink et al., 2005). When only the perimeters of structures were treated with 0.125% fipronil, 36% of the structures were still infested at 6 months postreatment (Waite et al., 2004). No area-wide effects on the colonies were observed. Similarly, Ripa et al. (2007) were unable to show any area-wide effects of perimeter treatments of fipronil. Fipronil is a phenylpyrazole insecticide that is extremely toxic to subterranean termites, with LD50 values ranging from 0.2 to 2 ng/termite (Ibrahim et al., 2003; Saran and Rust, 2007; Yamaoka and Tsunoda, 2007). It is slow-acting and nonrepellent with the maximal effects occurring 3–5 days after exposure (Remmen and Su, 2005a,b; Saran and Rust, 2007). Exposure to fipronil-treated deposits resulted in delayed toxicity (Hu, 2005; Saran and Rust, 2007). Horizontal transfer occurs only when donors are exposed to deposits with concentrations 5 ppm (Bagnères et al., 2009; Saran and Rust, 2007; Tsunoda, 2006), maximum transfer occurs within first 24 h (Saran and Rust, 2007), and contact and grooming are the primary mechanisms for transfer (Bagnères et al., 2009; Saran and Rust, 2007). Concentrations 100 ppm were needed to transfer fipronil among the Formosan subterranean termite, Coptotermes formosansus Shiraki (Shelton and Grace, 2003). In a laboratory study, the maximum mortality occurred within 1.5 m of the treated zone and the maximum effects 2.5 m from the treated zone (Saran and Rust, 2007). Su (2005) reported lethal transfer effects 5 m. The neonicotinoid imidacloprid is highly toxic to termite workers, with as little as 10.6 ng/termite providing 50% kill in 7 days (Rust and Saran, 2008). In addition to toxicity, sublethal effects have been reported with imidacloprid inhibiting feeding of workers of the eastern subterranean termite, Recticulitermes flavipes (Kollar), after exposure

Chapter | 5  The Changing Role of Insecticides in Structural Pest Control

(Ramakrishnan et al., 2000). Sublethal exposures to imidacloprid also decreased tunneling activity in R. virginicus (Thorne and Breisch, 2001) and locomotion (Haagsma and Rust, 2007). The deposits are nonrepellent and termites readily contact lethal treatments (Haagsma and Rust, 2007). Shelton and Grace (2003) found that exposures on 100 ppm deposits were necessary for transference. Imidacloprid is transferred between termites primarily by contact (Haagsma and Rust, 2007). Exposures to imidacloprid decreased locomotion within hours, probably minimizing the importance of horizontal transfer. This supported the finding of Osbrink et al. (2005) that there was no measurable horizontal transfer under field conditions. Chlorfenapyr is a slow-acting insecticide that is horizontally transferred by termites (Rust and Saran, 2006; Shelton et al., 2006) by mutual grooming and contact (Rust and Saran, 2006). Deposits ranging from 100 to 300 ppm were not repellent and as low as 10 ppm prevented tunneling (Rust and Saran, 2006). Termite movement was greatly hampered within 4 h after exposure to 50 ppm chlorfenapyr deposits, limiting the distance that termites might travel back to the nest. The neonicotinoid acetamiprid is extremely toxic to termites with as little as 0.02 and 0.04 ng/termite killing 50% of the western subterranean termite, Reticulitermes hesperus Banks, and R. flaviceps (Oshima), respectively (Mo et al., 2005; Rust and Saran, 2008). Barriers with as little as 1–8 ppm prevented termite tunneling. Contrary to the notion that all slow-acting neonicotinoids are nonrepellent, acetamiprid was repellent at all concentrations and acted more like a pyrethroid treatment in soil (Rust and Saran, 2008). With the advent of new chemistries such as the neonicotinoids, phenylpyrazoles, and pyrroles, changes in the ways that we pretreat and remedially treat structures for subterranean termites have occurred and more are likely in the future. With shorter residuals in the soil than provided by organochlorine, organophosphate, and pyrethroid insecticides, it may be necessary to re-treat structures more often as part of a contractual service. Since some of the newer termiticides are nonrepellent and actively kill foragers, the amount of termiticide applied and the extent of the coverage may be reduced. By targeting active infestations with repeated treatments, the industry may begin to move away from the concept of merely applying long-term protective barriers.

5.6  Baits and baiting Baits have been used to control structural insect pests for over 100 years. Typically, they have incorporated fastacting toxicants such as organochlorines, organophosphates, and carbamates (Rust, 1986, 2001). With the advent of slower-acting insecticides developed as baits for red imported fire ant control in the early 1980s (Klotz et al., 2008), bait development underwent a major revolution.

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A consequence of this research was the development of cockroach baits that changed the paradigm of German cockroach control (Reierson, 1995; Rust, 2001). Behavioral resistance or bait aversion was first described by Silverman and Bieman (1993) in a fieldcollected strain of German cockroach. The cockroaches avoided ingesting toxicant/diet mixtures containing d-glucose and this aversion was an autosomal incompletely dominant trait. Another field-collected strain showed avoidance to fructose, glucose, maltose, and sucrose and moderate levels of physiological resistance to abamectin and fipronil (Wang et al., 2004). Behavioral resistance to baits was weakly sex-linked, with females inheriting to a greater degree (Wang et al., 2006). These studies suggest that behavioral and physiological resistance will quickly develop to baits, especially if they are intensively used to control German cockroaches. Consequently, better IPM strategies that include improved bait bases and rotational schemes for toxicants will be necessary in the future. Secondary and tertiary mortality of cockroaches has been reported with baits containing indoxacarb (Buczkowski et al., 2008). Factors affecting the transfer included the freshness of excretions from donors, the presence of alternative food, and the length of time of contact between donors and recipients. Behavioral studies of German cockroaches and baits containing microencapsulated chlorpyrifos indicated that the amount consumed was affected by insect age and food nutrititional content (Jones and Raubenheimer, 2002). The toxicant affected the average meal duration and frequency in interaction with age and food nutrient effects. Nymphs that survived the initial exposure to bait were less likely to ingest lethal doses at a later time. The chitin synthesis inhibitor noviflumuron was active against nymphal B. germanica producing 99% kill of laboratory populations at 7 weeks (Wang and Bennett, 2006a). In simulated kitchens, mixed populations of B. germanica were reduced 96% by week 8. Indoxacarb baits were toxic and relatively nonrepellent, significantly reducing field populations of German cockroaches (Appel, 2003). In field studies, building-wide IPM programs including baiting were comparable to stand-alone baiting for 29 weeks (Wang and Bennett, 2006b). Baits and baiting technology are currently being marketed to control subterranean termites (Rust, 2001). The ideal bait toxicant for subterranean termites (1) must be slow acting, (2) must act as a nondeterrent, and (3) must not cause adverse effects when ingested at sublethal doses. Furthermore, its lethal time must be dose independent (Su and Scheffrahn, 1998). The uptake of hexaflumuron on worker R. hesperus peaked at about 280 ng at day 12 (Haagsma and Rust, 2005). Dead termites contained 113 ng of hexaflumuron suggesting that this is the minimal lethal dose required to kill workers. Hexaflumuron was quickly excreted by workers and the half-lives ranged from 2.1 to 4.7 days, but it was readily dispersed through

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the termite colony by trophallaxis. The uptake of the benzolphenylurea IGR noviflumuron by R. flavipes was similar to hexaflumuron, but the half-life in the termites was about 29 days compared with 8–9 days with hexaflumuron (Karr et al., 2004). The toxic dose of noviflumuron was two- to threefold less than hexaflumuron. Therefore, the faster activity of noviflumuron is in part explained by the longer retention in termite workers and its greater intrinsic activity. Another benzolphenylurea compound, bistrifluron (5000 ppm), provided slower toxicity to C. formosanus at 8 weeks (Kubota et al., 2006). Higher concentrations of bistrifluron caused some feeding repellency. A gel formulation of 2.15% hydramethylnon was tested against the western drywood termite, Incisitermes minor (Hagen), in small block tests (Indrayani et al., 2008). It was unclear whether termites had been killed by contact or ingestion of bait, but the results are interesting and warrant additional investigation. The development and requirements of a toxicant for an ant bait differ from those of termite baits because of the increased amount of trophallaxis and complete metamorphosis in ant colonies. Stringer et al. (1964) concluded that ant bait toxicants must exhibit delayed toxicity over at least 10- to 100-fold dosage range, not be repellent when added to bait, be readily transferred between ants, and kill the recipient. In addition, the delayed toxicity of the toxicant must not inhibit or prevent trophallaxis. Even though several toxicants such as hydramethylnon, sulfluramid, abamectin, and fipronil meet all the criteria, the development of acceptable bait bases for the various species of pest ants has been problematic. No single bait base is attractive to all pest species of ants. Consequently, special bait bases will be needed for different groups of ants. IGRs such as fenoxycarb, methoprene, and pyriproxyfen have been successfully used to control S. invicta, but little information is available on other pestiferous species such as Argentine ants, odorous house ants, and carpenter ants as reviewed by Klotz et al. (2008). Baits containing toxicants such as hydramethylnon provided delayed toxicity (15% mortality at 24 h; Stringer et al., 1964) of S. invicta workers. However, indoxacarb provided 15, 55, and 100% kill at 9, 24, and 48 h, respectively (Oi and Oi, 2006). Indoxacarb did not provide delayed toxicity and the results clearly suggest that our criteria for ant baits (based on Stringer et al., 1964) may need to be modified for pro-insecticides like indoxacarb.

5.7  Future directions The implementation of IPM programs for the control of structural pests and regulatory changes will increase in response to the public’s concerns over insecticide use, the so-called green pest control movement, and new challenges presented by invasive pests. For example, 24.5% of U.S. states require school districts to conduct IPM and 30.6%

recommend that IPM be conducted (Jones et al., 2007). This will certainly increase in the future. Insecticides will continue to play an extremely important role in urban pest management, but there will be increased emphasis on those that possess a real or perceived threat to urban waterways and indoor and outdoor environments. The types of insecticides applied and their use pattern will dramatically change over the next 10 years and more restrictions on the use of pyrethroids around structures are likely. The broadcast use of insecticides indoors and outdoors and barrier sprays outdoors will be phased out for more prescription treatments such as baits, crack and crevice applications, and traps. The popularity of baits will continue, stimulating new avenues of research dealing with the behavior and ecology of structural insect pests. Natural products and other chemicals such as essential and fragrance oils and detergents will also continue to be of interest. It remains to be seen if these products can be effectively incorporated into viable IPM programs. The consumer and public will demand more environmentally safer and effective treatments that use less insecticides and this will be the direction of urban pest management for the next decade.

Conclusion Reemerging pests and environmental challenges are changing the face of urban pest management. The reemergence of bedbugs and invasive pest species such as the red imported fire ant and Argentine ant pose special problems and challenges to the PMP. There is an increasing database and public awareness of the medical importance and role of cockroaches and other arthropods in inducing asthma and related health issues. Environmental monitoring suggests that outdoor perimeter treatments used to control ants may be an important source of insecticides in urban waterways. This development will likely result in major regulatory changes and use patterns of insecticides around urban structures and present challenges to the PMP and homeowner. Increasing awareness and environmental concerns of the public have resulted in a movement toward so-called green pest control. As a result, the use of botanicals, fragrance oils, and inorganic compounds has increased dramatically. Other new chemistries such as the neonicotinoids, phenylpyrazoles, and pyrroles have largely replaced the organophosphate and some traditional uses of pyrethroid insecticides. In the future, there will be a continued emphasis on low environmental impact strategies that incorporate the use of reduced-risk insecticides.

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Rust, M. K., and Saran, R. K. (2006). Toxicity, repellency, and transfer of chlorfenapyr against western subterranean termites (Isoptera: Rhinotermitidae). J. Econ. Entomol. 99, 864–872. Rust, M. K., and Saran, R. K. (2008). Toxicity, repellency, and effects of acetamiprid on western subterranean termite (Isoptera: Rhinotermitidae). J. Econ. Entomol. 101, 1360–1366. Saran, R. K., and Kamble, S. T. (2008). Concentration-dependent degradation of three termiticides in soil under laboratory conditions and their bioavailability to eastern subterranean termites (Isoptera: Rhinotermitidae). J. Econ. Entomol. 101, 1373–1383. Saran, R. K., and Rust, M. K. (2007). Toxicity, uptake, and transfer efficiency of fipronil in western subterranean termite (Isoptera: Rhinotermitidae). J. Econ. Entomol. 100, 495–508. Scharf, M. E., Ratliff, C. R., and Bennett, G. W. (2004). Impacts on residual insecticide barriers on perimeter-invading ants, with particular reference to the odorous house ant, Tapinoma sessile. J. Econ. Entomol. 97, 601–605. Sever, M. L., Arbes, S. J. Jr., Gore, J. C., Santangelo, R. G., Vaughn, B., Mitchell, H., Schal, C., and Zeldin, D. C. (2007). Cockroach allergen reduction by cockroach control alone in low-income urban homes: a randomized control trial. J. Allergy Clin. Immunol. 120, 849–855. Shelton, T. C., and Grace, J. K. (2003). Effect of exposure duration on transfer of nonrepellent termiticides among workers of Coptotermes formosanus Shiraki (Isoptera: Rhinotermitidae). J. Econ. Entomol. 96, 456–460. Shelton, T. C., Mulrooney, J. E., and Wagner, T. L. (2006). Transfer of chlorfenapyr among workers of Reticulitermes flavipes (Isoptera: Rhinotermitidae) in the laboratory. J. Econ. Entomol. 99, 886–892. Silverman, J., and Bieman, D. B. (1993). Glucose aversion in the German cockroach, Blattella germanica. J. Insect Physiol. 39, 925–933. Soeprono, A. M., and Rust, M. K. (2004a). The effect of delayed toxicity of chemical barriers to control Argentine ants (Hymenoptera: Formicidae). J. Econ. Entomol. 97, 2021–2028. Soeprono, A. M., and Rust, M. K. (2004b). Effect of horizontal transfer of barrier insecticides to control Argentine ants (Hymenoptera: Formicidae). J. Econ. Entomol. 97, 1675–1681. Stejskal, V. (2002). Inversion relationship between action threshold and economic/aesthetic injury level for control of urban and quarantine pests. J. Pest Sci. 75, 158–160. Stejskal, V. (2003). “Economic injury level” and preventive pest control. J. Pest Sci. 76, 170–172. Stejskal, V., Aulicky, R., and Pekar, S. (2007). Brief exposure of Blattella germanica (Blattodea) to insecticides formulated in various microcapsule size and applied on porous and non-porous surfaces. Pest Manage. Sci. 65, 93–98. Stout, D. M. II, and Leidy, R. B. (2000). A preliminary examination of the translocation of microencapsulated cyfluthrin following applications to the perimeter of residential dwellings. J. Environ. Sci. Health B35, 477–489. Stout, D. M., and Mason, M. A. (2003). The distribution of chlorpyrifos following a crack and crevice type application in the US EPA indoor air quality research house. Atmospheric Environ. 37, 39–40. Stringer, C. E. Jr., Lofgren, C. S., and Bartlett, F. J. (1964). Imported fire ant toxic bait studies: evaluation of toxicants. J. Econ. Entomol. 57, 233–249. Su, N.-Y. (2005). Response of the Formosan subterranean termites (Isoptera: Rhinotermitidae) to baits or nonrepellent termiticides in extended foraging arenas. J. Econ. Entomol. 98, 2143–2152. Su, N.-Y., and Scheffrahn, R. H. (1998). A review of subterranean termite control practices and prospects for integrated pest management. Int. Pest Manage. 3, 1–13.

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TDC Environmental. (2006). Pesticides in urban surface waters. Urban pesticide use trends annual report 2006. http://www.tdcenvironmental.com/UP3%20Use%20Report%202006.pdf TDC Environmental. (2008). Pesticides in urban surface waters. Urban pesticide use trends annual report 2008. http://www.up3project.org/ documents/Final_UP3_Use_Report_2008.pdf Temu, E. A., Minjas, J. N., Shiff, C. J., and Majala, A. (1999). Bedbug control by permethrin-impregnated bednets in Tanzania. Med. Vet. Entomol. 13, 457–459. Thorne, B. L., and Breisch, N. L. (2001). Effects of sublethal exposure to imidacloprid on subsequent behavior of subterranean termite Reticulitermes virginicus (Isoptera: Rhinotermitidae). J. Econ. Entomol. 94, 492–498. Toloza, A. C., Vassena, C., and Picollo, M. I. (2008). Ovicidal and adulticidal effects of monoterpenoids against permethrin-resistant human head lice, Pediculus humanus capitis. Med. Vet. Entomol. 22, 335–339. Tsunoda, K. (2006). Transfer of fipronil, a nonrepellent termiticide, from exposed workers of Coptotermes formosanus (Isoptera: Rhinotermitidae) to unexposed workers. Sociobiology 47, 563–575. Wagner, T., Shelton, T., Peterson, C., and Mulrooney, J. (2003). Putting termiticides to the test. Pest Control 72, 23–29. Waite, T. D., Gold, R. E., and Howell, H. N. (2004). Field studies of exterior-only applications with fipronil for post-construction control of interior populations of subterranean termites (Isoptera: Rhinotermitidae). Sociobiology 43, 221–229. Wang, C., and Bennett, G. W. (2006a). Comparative study of integrated pest management and baiting for German cockroach management in public housing. J. Econ. Entomol. 99, 879–885. Wang, C., and Bennett, G. W. (2006b). Efficacy of noviflumuron gel bait for control of the German cockroach, Blattella germanica (Dictyoptera: Blattellidae)—laboratory studies. Pest Manage. Sci. 62, 434–439. Wang, C., El-Nour, M. M. A., and Bennett, G. W. (2008). Survey of pest infestation, asthma, and allergy in low-income housing. J. Community Health 33, 31–39. Wang, C., Scharf, M. E., and Bennett, G. W. (2004). Behavioral and physiological resistance of the German cockroach to gel baits (Blattodea: Blattellidae). J. Econ. Entomol. 97, 2067–2072.

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Wang, C., Scharf, M. E., and Bennett, G. W. (2006). Genetic basis for resistance to gel baits, fipronil, and sugar-based attractants in German cockroaches (Dictyoptera: Blattellidae). J. Econ. Entomol. 99, 1761–1767. Wei, Y., Appel, A. G., Moar, W. J., and Liu, N. (2001). Pyrethroid resistance and cross-resistance in the German cockroach, Blattella germanica (L). Pest Manage. Sci. 57, 1055–1059. Weston, D. P., Holmes, R. W., and Lydy, M. J. (2009). Residential runoff as a source of pyrethroid pesticides to urban creeks. Environ. Pollution 157, 287–294. Weston, D. P., Holmes, R. W., You, J., and Lydy, M. J. (2005). Aquatic toxicity due to residential use of pyrethroid insecticides. Environ. Sci. Technol. 39, 9778–9784. Whyatt, R. M., Camann, D. E., Kinney, P. L., Reyes, A., Ramirez, J., Dietrich, J., Diaz, D., Holmes, D., and Perera, F. P. (2002). Residential pesticide use during pregnancy among a cohort of urban minority women. Environ. Health Perspect. 110, 507–514. Williams, G. M., Linker, M., Waldvogel, M. G., Leidy, R. B., and Schal, C. (2005). Comparison of conventional and integrated pest management programs in public schools. J. Econ. Entomol. 98, 1275–1283. Williams, M. K., Rundle, A., Holmes, D., Reyes, M., Hoepner, L. A., Barr, D. B., Camann, D. E., Perera, F. P., and Whyatt, R. M. (2008). Changes in pest infestation levels, self-reported pesticide use, and permethrin exposure during pregnancy after the 2000–2001 U.S. Environmental Protection Agency restriction of organophosphates. Environ. Health Perspect. 116, 1681–1688. Wiltz, B. A., Suiter, D. R., and Gardner, W. A. (2007). Deterrency and toxicity of essential oils to Argentine and red imported fire ants (Hymenoptera: Formicidae). J. Entomol. Sci. 42, 239–249. Yamaoka, R., and Tsunoda, K. (2007). Determination of the lethal dose of fipronil for workers of Coptotermes formosanus (Isoptera: Rhinotermitidae) in topical application. Sociobiology 50, 205–211. Yang, Y.-C., Lee, H.-S., Clark, J. M., and Ahn, Y.-J. (2004). Insecticidal activity of plant essential oils against Pediculus humanus capitis (Anoplura: Pediculidae). J. Med. Entomol. 41, 699–704. Zungoli, P. A., and Robinson, W. H. (1984). Feasibility of establishing an aesthetic injury level for German cockroach pest management program. Environ. Entomol. 13, 1453–1458.

Chapter 6

Vertebrate Pest Control Chemicals and Their Use in Urban and Rural Environments Rex E. Marsh1 and Terrell P. Salmon2 1 2

Wildlife, Fish and Conservation Biology, University of California, Davis, California University of California Cooperative Extension – San Diego County, San Diego, California

6.1  Introduction The scope and magnitude of vertebrate problems affect a wide and varied range of activities and health issues. Because of the breadth of the problems, the role of vertebrate pest management, in its broadest sense, is not easy to condense into a single chapter. Most people relate to specific vertebrate pest problems such as finding a house mouse in their kitchen or finding their car, parked beneath a tree, whitewashed with crow droppings, and have some idea as to how to resolve the problem. Few people recognize that the safety of the plane they are about to board may rely on a specialist in vertebrate pest control to see that the gulls are dispersed from the runway prior to takeoff, nor are they aware that an overpopulation of deer in many eastern states causes thousands of car accidents annually. There is also a general lack of knowledge that vertebrates contribute to higher prices for cereals, fruits, and nuts at the supermarket in order to compensate for crop losses resulting from bird and mammal feeding activities and that some means of reducing these losses are commonly employed. The significance of rats and salmonellosis, bats and rabies, deer and Lyme disease, deer mice and hantavirus, ground squirrels and plague may be common knowledge regionally, but the magnitude of disease problems associated with wild animals on a nationwide basis and the means used to minimize them are rarely discussed and are unfamiliar to most. Contrarily, rare attacks on humans by bears, coyotes, mountain lions, or alligators almost always receive exceptional news media attention. Although chemical agents are important, the majority of vertebrate pest problems are managed without the use of pesticides or chemicals (Hygnstrom et al., 1994; Salmon et al., 2006). The approaches and techniques used in vertebrate pest management are reviewed to illustrate and to Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

put into proper perspective the use of vertebrate pesticides compared to other methodologies. Some of the many types and situations of vertebrate pest problems are referenced, establishing the basis for the use of vertebrate pest control chemicals, both nontoxic repellents and lethal agents. The amount of vertebrate pesticides used is relatively minor compared with insecticides and herbicides. This is in part because the problems, when caught early, generally occur over relatively small acreages or involve a limited number of animals. Vertebrate pesticides are targeted to very specific habitats or to only the offending animals or populations. For example, when rodenticides are used efficiently for long-term population suppression, only small amounts of pesticides are needed. Control strategies are selected to be as target specific as possible, using the minimal amount of pesticide.

6.2  Vertebrate pests: what are they? Vertebrate pest control, animal damage control, wildlife damage control, and vertebrate pest management are all terms commonly used interchangeably. Vertebrate pests can be defined as any vertebrate, native or introduced, domestic or feral, that periodically or consistently has an adverse effect on human health and well-being or conflicts in some significant way with human activities or interests. Vertebrate pests are a diverse group of animals and include amphibians, reptiles, birds, and mammals (National Academy of Sciences, 1970). While fish are also vertebrates, they are not normally included when discussing vertebrate pests although there are significant fish pest problems. For most wildlife species, except possibly the introduced nonnative rats and mice, pigeons, house sparrows, and starlings, it is unfair to categorically characterize any native wildlife as a pest. Most are considered pests 271

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only in relatively few circumstances and may be neutral or highly desirable in most situations. For example, deer may be pests in some suburban areas because they cause automobile accidents and feed in gardens, yet in low or moderate numbers they may be treasured for their esthetic value on golf courses and in parks. Some enjoy feeding pigeons in the public square, whereas property owners in the immediate vicinity may view them as nothing more than messy pests. In rural regions, sport hunters prize elk as game species, but to a forester elk may be devastating in a newly established plantation. The sound of howling coyotes may be an enjoyable part of the camping experience, but to a sheep rancher, these sounds are nothing more than the prelude to lamb losses. Whether an animal is or is not considered a pest is often highly subjective and, as a result, vertebrate pest management can become highly emotional and controversial.

6.3  Management restrictions The management or control of vertebrate pests is generally much more restrictive and complicated than the control of insects, plant diseases, nematodes, or weeds. Unlike the plant and invertebrate groups, which can be controlled by any legal means by the property owner or manager, native wildlife and even some introduced species come under the jurisdiction of either the federal or the state governments. Legally, they are the property of all the people and therefore cannot be freely controlled by individuals. This means that property owners and pest management professionals must look to the laws and regulations governing wildlife for authority to control wild vertebrates that become pests. Wildlife agencies determine which species may or may not be controlled and sometimes dictate what control methods can be used. Federal and state wildlife agencies that permit or authorize the management or control of pest wildlife are, for the most part, conservation oriented and at the same time they also facilitate consumptive uses such as fishing, hunting, and fur trapping. The federal and state laws and regulations governing pest wildlife management or control are numerous and sometimes complicated, as are those regarding consumptive uses of wildlife.

6.4  Problems created by vertebrates Human/animal interactions or conflicts that become problematic are numerous and highly varied; entire volumes are devoted to the subject (Conover, 2002). For the purpose of this chapter, a representative selection of examples is provided to establish a basis for the need and use of vertebrate pest control agents. Vertebrate pest problems are categorized by the nature of the conflict. Table 6.1 provides a glimpse of the generalized problems relative to urban and

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suburban environments by listing the problems and then examples of offending animals (Marsh, 1986). A similar characterization is provided in Table 6.2 but is directed toward problems associated with agricultural production, forestry, and recreation and natural areas.

6.5  Nonlethal management without pesticides The majority of vertebrate pest problems are prevented or resolved without the use of pesticides and/or lethal controls. A variety of approaches are employed, including exclusion by fencing, netting, or rodent and bird proofing of buildings; habitat modification; selective crop cultural practices that include resistant crop varieties; and pestfrightening methods. Deer-proof fences are used to protect orchard crops and vineyards; predator fences offer protection to sheep, goats, and calves. Buildings are made rodent proof to block the entrance of rats and house mice, and bird netting is used to prevent pigeons, swallows, and house sparrows from nesting or roosting on buildings as well as to restrict entrance. Bird netting is also used to protect high-value crops such as grapes and blueberries. Habitat modification generally takes the form of removing the pest animal’s food and cover, making the area less attractive or even inhospitable. In urban settings, this may be accomplished by removing backyard debris and covering garbage cans to reduce rat populations. Vines and other plants close to buildings can be pruned or removed to help restrict access by rodents such as rats. Shade trees serving as major bird roosts may be removed or severely pruned to discourage crows, blackbirds, and starlings. Crop cultural practices, a form of habitat modification, are a major means of reducing specific vertebrate pest problems. Frequent and complete cultivation, as occurs with annual truck crops, discourages meadow voles, pocket gophers, and ground squirrels. Permanent cover crops kept closely mowed in orchards and vineyards reduces the population and damage caused by meadow voles, as does maintaining weed-free tree or vine rows. Weed-free fence rows and ditch banks are contributive to lower rodent populations, resulting in reduced damage to adjacent crops. Unfortunately, habitat changes affect all wildlife so targeting one species may lead to improving the habitat for another vertebrate pest. Habitat changes can also impact populations of other pest species such as predatory insects. To reduce losses from birds, varieties of sunflowers and sorghums can be selected for their resistance to blackbird damage. Alfalfa varieties have been evaluated for their resistance to pocket gopher damage and apple root stocks for their resistance to pine vole damage, giving the growers another means of reducing crop losses.

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Table 6.1  Nature of Select Vertebrate Pest Problems Occurring in Urban and Suburban Environments, along with Examples of the Pests Implicated and the Particular Issuesa Nature of problem

Examples

Animal-related health problems   Disease potential

Pigeon droppings and histoplasmosis; bats and potential rabies; rodent and bird contamination in restaurants and food-processing facilities

 Parasites and host-related arthropods

Pigeons and sparrows and mite infestations; deer mice and ticks; rats and fleas; raccoons and roundworms

  Animal conflicts and injuries

Rat bites to children, elderly, and pets; coyote, mountain lion, and alligator attacks on children; coyotes and mountain lions feeding on pet cats and dogs

  Vehicle collisions

Deer-caused automobile accidents; bird strikes contributing to aircraft accidents; deer and gulls making takeoffs unsafe

  Phobias to animals

Fear of bats, rodents, or snakes in the home; fear of almost all mammals or birds

Product and material losses   Food loss

Rats and house mice destroying human and animal food commodities in warehouses, residences, etc.

 Degradation or destruction of human-made structures and materials

Gnawing of doors, cabinets, electrical wiring by rats and house mice; woodpecker damage to houses; bird droppings defacing buildings; rodents destroying stored books, artwork, and manuscripts; raccoons ripping off shingles

 Communication and power outages and fires

Power outages caused by squirrels nesting or birds roosting in power substations; fires caused by rats and mice gnawing on electrical wires; pocket gophers gnawing through underground communication cables

  Vegetation destruction

Rabbit, pocket gopher, vole, and deer damage to landscape plantings, and flower and vegetable gardens

Environmental consequences   Noise disturbances

Starlings roosting on building ledges; rats and bats in attics or in walls; raccoons in the chimney

  Odor pollution

Skunks beneath houses; a dead rat within a wall; bird droppings in air-conditioning systems; roaming cats defecating and spraying in gardens

  Unsightly conditions

Bird droppings on sidewalks, boats, and vehicles; garbage scattered by raccoons, opossums, feral dogs, and bears; frogs and voles fallen into swimming pools

  Competition with pets

Rodents and raccoons stealing pet food and causing dogs to bark; birds diving at dogs and cats; rodents infesting outdoor aviaries

a

 These problems are in no way unique to urban environments and can also occur in rural environments.

Frightening methods involve everything from scarecrows to herding birds by aircraft. Propane exploders, eye balloons, reflective tape, predator effigies, recorded bird distress calls, roving patrols, and guard dogs are frequently used to frighten pests from an area or crop. The great advantage of all of these approaches is that they do not require killing any animal and are, therefore, acceptable to most people. Unfortunately, such approaches cannot be used to resolve all of the varied vertebrate pest problems. The expense of the approach may also be a limiting factor as it may not be economically sound. In a few instances, a control approach using sound-producing devices in cities or residential areas, such as propane exploders or

broadcast bird distress calls, is limited because of noise ordinances. Even the removal of a tree or construction of certain types of deer fences may be prohibited by building codes in some areas.

6.6  Population reduction without pesticides Several nonpesticide approaches are employed to kill offending animals or to reduce local pest populations. Hunting or shooting is used to take deer, mountain lions, bear, coyotes, and sometimes porcupines, woodchucks,

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Table 6.2  Characterization of Select Vertebrate Pest Problems More Commonly Associated with Agricultural Production, Forestry, and Outdoor Recreation, along with Examples of the Pests Implicated and the Particular Issues Nature of problem

Examples

Agricultural production losses   Fruit and nut crop damage

Crows and ground squirrels take almonds; tree squirrels feed on pecans; voles girdle apple trees; beavers cut down pear trees; rabbits clip young trees

  Forage and rangeland damage

Pocket gophers, prairie dogs, ground squirrels, kangaroo rats, and jackrabbits reduce production; elk and deer feed on haystacks

  Truck crop losses

Voles damage artichokes and tomatoes; pocket gophers devour carrots and other root crops; rat and pocket gopher burrows divert irrigation water

  Grain crop damage

Ground squirrels, jackrabbits, and deer reduce grain production; blackbirds consume maturing rice; rats feed on young rice plants; muskrat activity diverts water

  Field crop damage

Blackbirds feed on sunflowers; voles consume sugar beet; various rats feed on sugarcane; deer are a serious threat to soybeans

  Livestock-rearing losses

Coyotes and feral dogs prey on sheep and turkeys; mountain lions kill cattle and horses; eagles kill newborn kid goats and lambs; ravens peck the eyes out of newborn lambs; weasels and rats kill young poultry; rats and house mice often spread diseases and cause feed losses on dairy and swine farms and in poultry facilities

  Aquaculture–fish losses

Gulls, herons, and cormorants feed on fish; raccoons feed in commercial oyster beds

  Beekeeping losses

House mice and skunks destroy hives; bears in search of honey make a shambles of an apiary

Forestry production losses   Seedling losses

Pocket gophers sever roots; snow-shoe hare clip stems

  Young tree injury

Mountain beaver clip leaders; deer browse on new growth; deer scar bark by antler rubbing

  Maturing tree injury

Porcupines girdle tops; tree squirrels strip bark; bears claw the bark; woodpeckers reduce wood quality; beaver cut trees

Recreation and natural areas   Turf damage

Pocket gophers and moles damage lawns in cemeteries; Canada geese overrun soccer fields; rabbits and deer feed excessively in golf courses; skunks and raccoons dig up lawns in search of grubs; ground squirrels create hazardous burrows in schoolyard playing fields

  Natural vegetation damage

Large starling and blackbird roosts result in nitrogen-killed trees; beaver dams cause destructive flooding and death of trees; feral pigs root up natural vegetation

prairie dogs, and ground squirrels. Trapping is a major vertebrate pest management tool for reducing pest rodent and rabbit populations and for eliminating livestock predators. The control of rats and house mice relies heavily on trapping, which is an especially useful control method in residences. In urban and suburban situations, live-catch-type cage traps are used in some situations involving birds such as pigeons, house sparrows, and starlings, and for nuisance animal control involving tree squirrels, raccoons, skunks, opossum, and armadillos. Trapped animals may be relocated, depending on the species and state or federal laws, or euthanized. However, relocation is seldom recommended as a pest management strategy because of the potential of creating a pest problem elsewhere.

6.7  Pesticides: repellents versus lethal agents Vertebrate pesticides are represented by two major groups of agents, the lethal or toxic compounds selected or developed for killing the pest animal and the chemical repellents that are used to discourage pests from feeding or gnawing on some item or to repel them from a particular area. Both groups are registered as pesticides by the U.S. Environmental Protection Agency (EPA) under the Federal Insecticide, Fungicide and Rodenticide Act (FIFRA). Most perceive pesticides as being toxic, or at least toxic to the species for which they are intended. In the case of vertebrate pesticides, this results in a grossly distorted view of use practices. This distortion becomes apparent when

Chapter | 6  Vertebrate Pest Control Chemicals and Their Use in Urban and Rural Environments

compilations are made relative to the amounts of vertebrate pesticides used, as more often than not, both repellents and toxic agents are lumped together, establishing an unrealistic figure, which is then frequently used to address potential pesticide hazards and environmental concerns. To refer to repellents as pesticides lumps them with other materials that are designed or used to kill. Except for mosquitoes, repellents play a relatively insignificant role in insect or other pest management. In vertebrate pest management, however, a great emphasis is placed on the use of repellents, although their effectiveness is limited in many situations. A larger number and a greater variety of repellent products are marketed for controlling vertebrate pests than all other pests combined. By some calculations, the quantity of active ingredients marketed in the United States as vertebrate pest repellents far exceeds those sold as lethal control agents. The vast majority of repellents are nontoxic or essentially nontoxic as used and present few environmental problems. However, there are a few vertebrate repellent agents, such as naphthalene, that are toxic if ingested in sufficient amounts but not toxic to the target animal if used in accordance with the label directions. Many repellents are marketed for vertebrate pests, such as deer, cats, dogs, and a variety of birds, for which no lethal or toxic agent can or would be used. Although many repellents play an important role, they are not an effective solution for many problems. Most “taste” or “odor” repellents give limited and, at best, relatively short-term results. With repellents, results are often inconsistent and unpredictable. Lethal pesticides, on the other hand, generally give reasonably predictable results and are mostly directed at pests for which no other effective or economical remedy exists, at least under certain circumstances. The control of rats in blighted urban areas is a good example, as are meadow vole, pocket gopher, and ground squirrel control in agricultural and forestry situations. Lethal agents, including toxic baits, burrow fumigants, and tracking powders, are for the most part used against pest rodent and rabbit species. Baits and other toxic materials find limited use for some bird problems. Nonfood toxicant delivery methods such as M-44 injectors and toxic collars play a role in controlling livestock predators, especially in the Midwest and far western area of the United States. Depending on which pest is being controlled, lethal agents are referred to as rodenticides, avicides, or predacides. The characteristics and use practices for both repellents and lethal agents are addressed in greater detail under separate headings.

6.8  Repellents Chemical repellents are used in an attempt to manage certain vertebrate pests such as deer, rabbits, rats, dogs, snakes, bats, and a wide variety of others. In fact, there are few vertebrate pests to which chemical repellents have not

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been directed although most efforts are not very successful. Repellents are growing in popularity as a means of reducing damage without harming the pest. They are advertised to protect plantings or materials or to cause the animals to move from the location where they are causing a problem. Unfortunately many of the advertised claims have not been substantiated with rigorous laboratory or field testing. Chemical repellents are categorized in different ways; one of the simplest methods is to group them according to their perceived action, that is, “taste” and “odor.” “Taste” repellents are applied to plants or other objects to deter feeding or gnawing, whereas “odor” repellents are odiferous materials that repel the offending animals from the vicinity where applied (i.e., an “area repellent”). Although such a categorization of repellents is easy to understand, it is an oversimplification of behavioral results. Categorizing repellents based on more specific psychological and physiological consequences makes it possible to more accurately describe the potential mode or modes of action of active ingredients of repellents and their application (Mason, 1998). Using a more scientifically based classification, repellents are categorized as follows: 1. Predator-like odors (instills fear in herbivores) 2. Intense novel odors (fear of the new; neophobia) 3. Irritants (causes pain, making food items unpalatable or an area less hospitable) 4. Distasteful substances (alters the palatability of food items) 5. Taste-aversive conditioners (causes illness after ingestion and is rejected thereafter) No repellent classification is definitive or mutually exclusive, and some individual active ingredients can be placed in one or more groups. The first two, predator-like and intense novel odors, may be considered area repellents as the pest often does not have to taste or ingest the material. Predator odors or predatorlike odors generally contain sulfur compounds. Sulfur compounds are a byproduct of meat eaters and hence provide some short-term innate fear in herbivores that are frequently preyed upon. Deer Away Big Game Repellent and Hinder are examples of this type of product, as are a variety of predator urines sold commercially as deer and rabbit repellents. Novel odors are often sold as cat and dog repellents but are also marketed as area repellents for deer. These may include various volatile citrus, mint, or other plant oils that have fairly intense odors. Strongly scented soaps, aromatic garlic powder, and naphthalene also fall into this category. The effectiveness of novel odors from volatile substances is generally short-term as the animal readily habituates to them. Some animals may not be in the least affected. Exposure to volatiles from, or ingestion of, irritating compounds causes sensory pain. Such irritants are more than just an objectionable taste or smell because they stimulate the trigeminal pain receptors located in the mucous

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membranes of the eyes, mouth, and nose. Irritants are effective for some animals but may have little effect on others, especially birds; methyl anthranilate, the grapelike flavor found in ReJexIt, is an effective irritant for some birds but ineffective for mammals. Animals normally do not habituate to these repellents as long as the repellent remains at the effective concentration. Distasteful substances, such as castor oil, are essentially objectionable-tasting materials that are normally sprayed or dusted on plants to reduce their palatability to herbivorous animals such as deer, rabbits, and squirrels or applied to trash bags to prevent their being torn open by raccoons or dogs. Bitrex and denatonium saccharide are examples of bitter taste repellents that may influence palatability but have no latent physiological effects when consumed. Unfortunately, such bitter compounds have been found relatively ineffective against deer, rabbits, or other herbivores. Depending on their hunger drive, even raccoons, opossum, and dogs may habituate to these materials over time. Taste-aversive conditioners are chemical agents that, when ingested, are followed shortly by an illness and are then avoided by the animal after one or more exposures. This effect is referred to as a “conditioned aversion” or “learned taste avoidance.” It is the natural behavioral trait that protects animals from consuming toxic or harmful plants. Learned avoidance can result from a single exposure, especially when the illness is significant and has a rapid onset, and the taste of the material is new or novel. Methiocarb (Mesurol), thiram, and ziram are examples of repellents that result in such gastrointestinal distress when ingested. Some agents may exhibit more than one mode of action; in other instances, combining materials may result in superior repellency. The addition of a visual cue such as a colored dye or pigment as part of the repellent may further enhance their effectiveness. Whether any repellent will be effective depends on a number of factors, including the population density of the pest animal, the tenacity of a pest species to stay in the area, the availability of alternative food resources, and the desirability of the treated food to the offending animals. The plant growth characteristics and weather also affect results because new plant growth may lack treatment and rain may wash off or otherwise diminish the existing treatment. Repellents can be registered with the U.S. EPA without significant efficacy data; however, the U.S. EPA does require evidence that such a repellent is environmentally safe and nontoxic. Those repellents that are mildly toxic or not environmentally benign must undergo a more rigorous registration process. This lax regulatory requirement has led to a proliferation of marketed repellent products that lack any meaningful supportive laboratory or field efficacy data. The consumer is at a great disadvantage because there is little or no way of knowing if the repellent will be effective; unfortunately, many are not. Selecting a repellent

Hayes’ Handbook of Pesticide Toxicology

marketed by a reputable company works to the benefit of the consumer but is in no way a guarantee of effectiveness. Fortunately, there are a handful of relatively effective repellents available for which extensive efficacy data exist, and it is most worthwhile to seek out this information from appropriate wildlife authorities. The effectiveness of even the best of the available repellents may vary greatly, depending on the pest species, mode or modes of action, and the circumstances of use. Results are often inconsistent and failures or near failures are common. Use patterns vary widely, depending on the situation and pest implicated. Repellents are applied to plants or planted areas as sprays, dusts, or granules. Volatile odorproducing repellents are also enclosed in sachets, small mesh bags, or in specialized clip-on containers and placed or hung in open or confined areas to discourage the pest. Common uses include protecting gardens and forest plantings from deer, keeping birds and rodents from feeding on newly planted seed, repelling snakes from gardens and campsites and bats from attics, preventing birds from feeding on ripening fruit, and discouraging waterfowl from lawn areas. For reasons of food safety, repellents applied to edible crops must be registered for that purpose; most are not. A sixth category is represented by a group of repellents that is in no way associated with taste, smell, or illness, but rather relies on the sense of touch. These are sticky or tacky (tactile) repellents used mostly to discourage pest birds from perching and roosting sites such as building ledges, overhead beams, and pipes. When birds attempt to alight on treated surfaces, the tackiness underfoot spooks the birds and they move elsewhere. Tactile repellents are mostly formulated from polybutenes and are nontoxic. Such repellents are designed to be just sticky enough to discourage landing but not sticky enough to entrap the birds. Temperature extremes negatively affect efficacy, as does an accumulation of dust or debris on the surface of the applied repellents. Therefore, if the bird problem persists, frequent reapplication may be necessary. These repellents are formulated as liquids or semiliquids that are applied as sprays or formulated as a paste and applied with a caulking gun. A wide strip of adhesivebacked tape is sometimes laid down first and the repellent applied on top; this facilitates the removal of the repellent once results have been achieved. For the most part, tactile repellents are used to keep birds such as pigeons, sparrows, and starlings from perching on or in structures. Occasionally, they are applied to trees to prevent bird roosting. At least one product is registered for use to repel tree squirrels from climbing on buildings and/or from climbing bird feeder poles. Sticky-type repellents are mostly used by the structural pest management industry. Use by the general public

Chapter | 6  Vertebrate Pest Control Chemicals and Their Use in Urban and Rural Environments

is relatively rare because of their cost and the difficulty of application. Sticky-type repellent use for birds generally provides moderate to good results when applied according to directions. Frequent reapplication is often necessary.

6.9  Immobilizing agents One approach to dealing with bird problems is to use the immobilizing agent a-chloralose (Hygnstrom et al., 1994). This material, when consumed in bait, immobilizes the bird so it can be picked up and relocated. a-chloralose, in low doses, depresses the cortical centers of the brain. Waterfowl fed about 30 mg of a-chloralose per kilogram of body weight become comatose in 20–90 min with full recovery in 4–24 h. a-chloralose is best suited for capturing individual or small groups of problem waterfowl in situations or at times when other methods are not safe or practical. The U.S. Food and Drug Administration (FDA) has approved a-chloralose as an immobilizing agent for the U.S. Department of Agriculture–Animal and Plant Health Inspection Service–Animal Damage Control (USDA– APHIS–ADC) program to use in the capture of waterfowl, coots, and pigeons. More recently, approval has been granted for ravens and sandhill cranes. This use is granted exclusively to ADC under a continuing Investigational New Animal Drug (INAD) application. a-chloralose may only be used by ADC employees or biologists of other state or federal wildlife management agencies that have been certified in its use, or persons under their direct supervision.

6.10  Lethal vertebrate pesticides Toxic vertebrate pesticides are inherently hazardous because they are developed or selected for their ability to kill pest birds and mammals. They are generally considered more toxic than other kinds of pesticides since their target species is biologically more similar to humans. With very few exceptions, toxicants lethal to rodents, carnivores, and birds are also toxic to humans to the approximate same degree. Because of this potential hazard to people and domestic and other nontarget animals, use practices and delivery methods are designed to minimize exposure to nontarget species. Also, the availability of many of the most lethal agents is restricted to those individuals properly trained and certified in their application. It is also why federal and state pesticide regulators require such extensive product data for every lethal vertebrate pesticide. In the absence of satisfactory first aid treatments or antidotes and sufficient means of mitigating significant potential hazards to nontarget species, the product will likely not be registered for use. The careless or negligent use of some lethal agents has resulted in the accidental poisoning of pets, domestic stock, and, on very rare occasions, humans. Accidental

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exposure of dogs to rodenticides is the most prevalent of nontarget hazards. Children, especially those younger than the age of 3, sometimes get into rodent baits put out to control commensal rodents (i.e., rats or house mice). A small child seen playing with rodent bait, whether or not any has been seen to be ingested, generally results in a trip to the emergency room. Fortunately, however, considering the amounts of rodenticides used nationwide, there have been very few accidental human deaths resulting from the use of rodent baits. Criminal poisonings and suicides in the human population using such products are rare but always possible. Malicious poisoning of pets and even livestock or wildlife unfortunately does sometimes occur. The rarity of human fatalities can, for the most part, be attributed to the development and widespread use of anticoagulant rodenticides. Anticoagulant rodenticides are sufficiently slow acting and have detectable and identifiable symptoms that permit the antidote to be administered in a timely manner. Recovery from ingestion is generally complete, with no lasting or permanent damage. It is estimated that 95–98% of all rat and house mouse control conducted with poison baits in urban and suburban situations, including residences, is accomplished with the use of one or more of the available anticoagulant rodenticides. The question sometimes arises as to why lethal rodenticides are needed in urban and suburban situations for commensal rodent control as there is some risk to children and pets. The role that rodents play in the transmission of human diseases, many of which are fatal, far outweighs the very minor and controllable risk to children from accidental rodenticide exposure. Leptospirosis, murine typhus, plague, hantavirus, rat-bite fever, and salmonellosis, to name a few, are important parts of the equation when evaluating the benefits of rodenticides. Rodenticides have the ability to quickly and efficiently bring a large rodent infestation under control and are far more costeffective than other means of eliminating an existing population. Public health concerns generally tip the balance toward the retention and continued judicious use of rodenticides. Rodenticides, avicides, and predacides used in rural settings to protect forestry, crop, and livestock production and for rural public health issues are very rarely implicated in human exposure but are sometimes implicated in incidental wildlife losses. Lethal products currently registered have undergone close scrutiny, with restrictions and use practices designed to minimize potential nontarget losses. Again, considering the quantity of pesticides used nationwide, especially rodenticides, nontarget wildlife losses, with the current use restrictions in place, are negligible when put in proper perspective.

6.10.1  Poison Rodent Baits The great majority of lethal (toxic) pesticides employed in vertebrate pest control are used in bait form and kill the

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animal following ingestion. Most of these are used to kill pest rodents or rabbits. The largest use for rodenticide baits is for the control of house mice and Norway and roof rats in urban and suburban environments (Corrigan, 2001). In rural settings, voles, pocket gophers, ground squirrels, prairie dogs, cotton rats, kangaroo rats, wood rats, deer mice, muskrats, and nutria are representative of major agricultural or forestry rodent pests controlled with the use of poison baits. Poison baits are used in limited amounts for rabbit control. For illustrative purposes, rodenticides used in baits are separated into two groups, the anticoagulants and all others. Although all anticoagulant rodenticides have the same mode of action, the modes of action of those in the second group vary. The first-generation anticoagulants used in the United States include chlorophacinone, diphacinone, pindone, and warfarin. The second-generation series includes brodifacoum, bromadiolone, and difethialone. The nonanticoagulant group includes bromethalin, cholecalciferol, strychnine, and zinc phosphide. The rodenticides and major pest species for which they may be used are provided in Table 6.3. The second-generation anticoagulants are considerably more toxic than the first-generation materials and were developed specifically to combat a growing problem of genetic resistance to the first-generation anticoagulants occurring in rats and house mice in various locations across the country and world. Because of this higher toxicity, there are greater limits placed on where they may be used and who can apply them. This is especially true with the U.S. EPA’s new rules. By the year 2011, the second-generation anticoagulant baits will be restricted mostly to structural pest management professionals for use in and around buildings. They have never been permitted for use in the control of field rodents or rabbits in the United States.

6.10.1.1  Rodent Bait Preparation Baits are formulated by incorporating food items highly preferred by the target species with a sufficient amount of toxicant to result in death following ingestion. Bait may contain as much as 2% active ingredient, as is the case with the rodenticide zinc phosphide, or as little as 25 ppm of the anticoagulant difethialone. Anticoagulant baits are normally formulated in a range of 0.025–0.0025% active ingredient, depending on the specific anticoagulant and targeted rodent species. Baits formulated for rodents are generally cereal based and made of grains such as oats, wheat, barley, or corn, or a combination thereof. Formulations may also contain other ingredients such as adherents to bond the toxicant to the grain particles. Coloring agents often are used to assist in identifying the bait as being toxic and to aid in safeguarding nontarget species by repelling seed-eating birds from

Hayes’ Handbook of Pesticide Toxicology

consuming field-applied rodent baits. Bait formulated for commensal rats and mice often contains about 5% sugar or comparable sweetening agent and about 5% vegetable oil. These two ingredients enhance bait acceptance and consumption, resulting in an increased bait efficacy. Other inert ingredients such as oils and dyes can effect bait consumption as well (Salmon and Dochtermann, 2006). The U.S. EPA requires that baits for most rodents meet a relatively high degree of efficacy and have a reasonably good shelf-life. As a safety measure to avoid accidental ingestion, especially by young children, an extremely bitter agent called Bitrex is incorporated in small amounts into many of the baits prepared for commensal rat and house mouse control. In theory, this additive will make the bait so bitter that a child will spit it out before actual ingestion. Claims are also made that the bitter agent will offer some protection to pets as well. The material does have some adverse effect on rodent acceptance, but it may not appreciably impact control. Cereal-based bait is by far the most common rodenticide. It may, however, be prepared in different forms; the grain may be left whole, cracked, or rolled flat. It may also be coarsely or finely ground, and it is frequently pelletized or extruded into particles several times the size of a grain of wheat or kernel of corn. There are advantages to the various types of baits, depending on where and how they are applied and for which rodent they are intended. There are also baits prepared in a solid paraffin matrix. These generally consist of the anticoagulant in a mixture of ground grains that are cast or extruded into solid blocks ranging in size from about 2 oz to 1 lb. Solid blocks of bait were originally made in a paraffin matrix and intended for use in damp and moist conditions, for example, Norway rat control in sewers. Paraffinized cereal-based baits are more weather resistant and less apt to mold or mildew over time, making them acceptable to the rodents for a longer period. Such bait blocks are sufficient to provide the multiple daily feedings necessary to result in adequate control, especially with the first-generation anticoagulant-type rodenticides. Because of their convenience, efficacy, and safety, paraffinized rodent baits are currently extensively used in rat and mouse control in all types of situations. Rodent baits are marketed in several ways. Many baits registered for commensal rodent control can be purchased as over-the-counter products by the general public. Farmers may purchase rodent bait products for use around farm buildings; however, to control field rodents, the farmer or an employee may have to be a certified applicator in order to purchase and use specific rodent baits such as those containing strychnine or zinc phosphide. To apply some rodent baits classified as “Restricted Use Pesticides,” a permit is required and certain restrictions may apply. The target pest and intended use of the rodenticides must be specified on the product label.

Chlorophacinone

Ground squirrels

Prairie dogs

Cotton rats

Kangaroo rats

Wood rats

Deer mice

Muskrats

Nutria

Rabbits

Bromadiolone

a

Pocket gophers

Brodifacoum

Voles

a

Roof rats

Anticoagulants

Norway rats

Major pests

House mice

Rodenticide

—

—

—

—

—

—

—

—

—

—

—

—

—

—







—

—

—

—

—

—

—

—

—

—

—







—

—

—

—

—

—

—

—

—

—

—













—

—

—

—

—



—



a







—

—

—

—

—

—

—

—

—

—

—

Diphacinone













—

—

—

—

—



—



Pindone







—

—

—

—

—

—

—

—

—

—

—

Warfarin









—

—

—

—

—

—

—

—

—

—

Nonanticoagulants

—

—

—

—

—

—

—

—

—

—

—

—

—

—

Bromethalin







—

—

—

—

—

—

—

—

—

—

—

Cholecalciferol







—

—

—

—

—

—

—

—

—

—

—

Strychnine

—

—

—

—

x

—

—

—

—

—

—

—

—

—

Zinc phosphide



























—

Difethialone

Chapter | 6  Vertebrate Pest Control Chemicals and Their Use in Urban and Rural Environments

Table 6.3  Pesticides Commonly Used as Baits to Kill Major Pest Rodents and Rabbits

a

 Second-generation anticoagulants. x, indicates that there is a product(s) containing that ai that is used for the pest listed; —, indicates that no product with that ai is registered for that pest.

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Not all baits are cereal based and made of grain; perishabletype baits are also prepared using chunks or cubes of fresh fruit or vegetables. Muskrat and nutria populations that must be reduced may be controlled with zinc phosphide baits prepared with cut-up chunks of apples, sweet potatoes, or carrots. These are prepared by coating the chunks with a vegetable oil and then blending them with the appropriate amount of zinc phosphide concentrate. Such perishable fruit or vegetable baits must be prepared fresh and applied shortly thereafter, prior to deterioration. Such baits are relatively expensive to prepare and therefore are only used for the control of those pest species that have little fondness for cereals. When preparing these types of baits, follow the label instructions carefully. Liquid baits may be an appropriate alternative to food baits for some situations. Because most animals are attracted to water and utilize it, liquid baits are not very selective for the target species. For this and other reasons, their use is limited to the structural pest management professionals for the control of commensal rats and house mice in select situations such as warehouses and manufacturing facilities where cereal-based baits may not be well accepted for various reasons. Water or liquid baits are prepared with water-soluble sodium salts of anticoagulants such as diphacinone, pindone, or warfarin.

6.10.1.2  Methods of Bait Use In practice, toxic rodent baits are used or applied in various ways, depending on the pest and the situation or crop. In urban environments, rodent baits are placed as close as possible to the location where rodent sign or evidence of their activities is present. For safety concerns, baits are placed in locations inaccessible to children and pets or placed in tamper-resistant bait stations. Tamper-resistant bait stations are enclosed boxes usually made of metal or sturdy plastic and are of a sufficient size to permit rats or mice to enter through one of two small holes located on opposite sides of the station. Not only is the toxic bait placed within these stations protected from access by larger nontarget species but also the rodents are provided a secluded location to feed on the bait. The amount of bait placed in each station may vary from about 6 oz for house mice to several pounds for rats. The criteria for a tamperresistant bait station are established by the U.S. EPA and require that its design be such that a young child cannot reach into the bait station through the holes to touch the reservoir of bait. To be tamper resistant, the stations must, by some means, be anchored to the floor or ground so that bait cannot be shaken out, and the lid must be secured so that a child cannot gain access. Structural pest management professionals routinely use tamper-resistant bait stations when controlling commensal rats and mice. The latest U.S. EPA rules regarding consumer overthe-counter sales will limit the product to the first-generation

Hayes’ Handbook of Pesticide Toxicology

anticoagulant rodenticides. Each retail unit must be 1 pound or less and be inside a preloaded bait station. Loose bait forms, such as ground cereal meals, whole grains, or pellets, are prohibited. Only solid-type bait blocks (e.g., paraffinized blocks) will be available as commensal rodent bait products sold to the general public. These new rulings, to be fully implemented by 2011, may have some far-reaching implications for commensal rodent management—with decreasing control effectiveness and an increasing potential for greater rodent resistance to firstgeneration anticoagulants. Toxic water or liquid baits are often presented to the rodents in shallow containers placed within a bait station. Liquid rodent baits are generally used within industrialtype buildings such as warehouses and other enclosed facilities, providing all other animals are excluded. Water baits are sometimes offered in reservoir-type chick fountains, which hold a larger supply of liquid bait. Water baits are very rarely used even by professionals. They are not advantageous in most situations; they are time-consuming to prepare, and evaporation and deterioration are relatively rapid. Ingredients to prepare water baits are not marketed to the public. In rural settings, anticoagulant-filled bait stations are commonly used around farm buildings for the control of house mice, deer mice, and rats because they prevent pet and livestock access to the poison bait. In agricultural crops, bait stations, although not necessarily of the tamper-resistant type, are frequently used when anticoagulant-type baits are applied for ground squirrels, Norway and roof rats, and muskrat. Bait stations can be designed to exclude particular species of interest. For example, when baiting California ground squirrels, specially designed bait stations have been devised to exclude the endangered San Joaquin kit fox as well as endangered kangaroo rats. Other designs exclude threatened deer mice when baiting for roof rats. Baits for some agricultural pests are placed within the burrows, as is done for pocket gophers, or in the burrow entrances, as is sometimes done for Norway rats and ground squirrels. As previously mentioned, perishable baits of cut-up apples, carrots, or sweet potatoes find limited use. When used for the control of muskrats, they are placed in floating bait stations; when used for nutria, they are offered on floating rafts anchored away from the bank. Perishabletype baits prepared with zinc phosphide fall under the “Restricted Use” category and require some expertise to prepare and safely use. Native muskrats are not a serious pest problem in most regions, and the introduced nutria is principally limited in distribution to the southern states, especially Texas and Louisiana. A common method of bait application used in agriculture is called “spot baiting” in which a small amount of bait is scattered on bare ground near the burrow entrance, as is done with zinc phosphide baits for prairie dog or

Chapter | 6  Vertebrate Pest Control Chemicals and Their Use in Urban and Rural Environments

ground squirrel control, or placed in the burrow openings or trails of voles. Baiting of voles in apple orchards, vineyards, sugar beet fields, and on noncropland is commonly conducted by broadcasting, using some type of tractor or ATV-mounted power seeder calibrated to deliver the precise amount of bait per acre. Baits for ground squirrels are also sometimes applied by broadcasting. In some instances, the bait for vole control may be broadcast by airplane, when the acreages are large or when the ground may be too wet and soft to accommodate a vehicle-mounted broadcaster. Aircraft are also used to broadcast baits for the control of various rats that damage sugarcane, including native cotton rats, rice rats, and Florida water rats, as well as the introduced Norway and roof rats. Rat control in sugarcane grown in Hawaii includes the Polynesian rat, which is not found on the mainland. Maturing sugarcane is tall and dense, making other types of bait application difficult and impractical.

6.10.2  Fumigants Fumigants are either toxic gases or substances that produce toxic gases that are lethal when inhaled. In vertebrate pest control, fumigants are principally used to control rodents in one of two ways, as a building or transportation vehicle fumigant or as a burrow fumigant. Fumigants have many advantages over other control methods because they do not require any particular behavior or action on the part of the target animal. Fumigation of buildings, rail cars, etc. is often conducted for insect control and, depending on the fumigant, the process can also provide rat and house mouse control. Fumigation of buildings specifically for rodent control is sometimes conducted but it is generally prohibitively expensive. Building fumigation can only be conducted by licensed pest management professionals under a strict set of regulations. Burrow fumigants are used outdoors against a wide variety of burrowing rodents, including Norway rats, chipmunks, ground squirrels, prairie dogs, woodchucks (marmots), voles, and pocket gophers. Fumigants are also used, to a limited extent, as burrow or den fumigants to control certain carnivorous species such as coyotes, foxes, and skunks. Some are registered for use against moles; however, moles are not easily controlled with burrow fumigants. There are two fumigants that are commonly used in vertebrate pest control, aluminum phosphide and ignitable gas cartridges. Aluminum phosphide, a “Restricted Use Pesticide” to be used only by certified applicators, comes in tablet or pellet form. When the prescribed number of tablets or pellets is placed well within the burrow or den, they react with the soil or atmospheric moisture to produce lethal phosphine gas. The burrows or dens are sealed off with soil immediately following treatment to retain as

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much of the toxic gas as is possible and for as long as possible. If there are multiple entrances to a burrow or den, each entrance must be sealed so the toxic gas remains in the burrow. The other and more commonly used fumigant is the ignitable gas cartridge, which is sold over-the-counter to the public. There are several manufacturers of these cartridges, but all generally contain two basic ingredients, sodium nitrate and charcoal, combined with smaller amounts of active or inert ingredients. The formulated ingredients are compressed into a cardboard tube with a fuse inserted in one end. When ignited, they produce a toxic suffocating smoke that is lethal to animals in a confined space. To use, the cartridges are placed in the burrow or den entrance and the fuse is lit. Once lit, the cartridge is pushed deep into the burrow or den with a shovel handle and the opening is sealed off with a soil plug and tamped tightly to retain the smoke. When used in accordance with the directions, gas cartridges present little hazard to the user. Fumigants must remain in the burrow in sufficient concentration and time to be effective. Soil type and moisture level can impact the effectiveness of this control method. Dry and loose, rocky or sandy soil is less likely to maintain the gas concentration needed to be effective. Also, many fumigants, especially the gas cartridge, present a fire hazard and should not be used when dry grass or other flammable materials are present. Since rodent burrows can go beneath buildings, fumigants should not be used in close vicinity to buildings. There are machines that use a fumigation-type approach in an attempt to control burrowing rodents including pocket gophers and ground squirrels. These are the burrow exploders that inject oxygen and propane into the burrow and ignite the mixture, causing a significant explosion. There is very little reported evidence that these devises are effective in controlling ground squirrels or other burrowing rodents.

6.10.3  Tracking Powders Toxic tracking powders for commensal rodent control are applied in a thin layer on a solid surface where rats or house mice travel. When a rodent runs over a patch of toxic powder, the fine particles adhere to its feet and fur. Because rodents characteristically groom themselves by licking their paws and fur, sufficient toxicant is ingested to be lethal. Ingestion is the means of exposure to these pesticides, as skin adsorption and inhalation are negligible. Tracking powders, sometimes referred to as grooming toxicants, are frequently formulated in fine clay at active ingredient concentrations substantially greater than normal rodent baits. This is because “tracking and grooming” is not a highly efficient method of delivering a toxicant to the target species. Zinc phosphide and the anticoagulants

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(chlorophacinone and diphacinone) are the toxicants commonly used in formulating tracking powders. Tracking powders are applied with some type of duster over a rectangular area of a few inches wide, 12–18 in. in length, and about 1/16 in depth. The treated spots are commonly referred to as “tracking patches” and are usually placed along walls where signs and other evidence suggest a rodent travel way. Tracking powders can also be placed in specially made tubes with open ends, which allow the rodent to run through them, or they can be placed in shallow metal trays to facilitate easy cleanup. Tracking powders are, in addition, blown into wall voids where rodents are known to thrive. When used outdoors, tracking powders are not as effective; however, they are sometimes placed in the burrow entrance of Norway rats. They are usually ineffective if they become wet. These toxic powders are more effective on house mice than on Norway or roof rats because, in proportion to their size, more toxic powder adheres to house mice than to rats. Mice also spend more time grooming than do rats. They are not recommended for use in food-processing facilities or other critical areas where toxicant might be tracked to living spaces or food commodities. Tracking powders are “Restricted Use Pesticides” and not sold to the public. They are principally used by structural pest management professionals.

6.10.4  Contraceptives Research on reproductive control in wildlife management (e.g., deer, wild horses, Canada geese, coyotes, and various rodents) has been ongoing for many years, especially at the USDA National Wildlife Research Center, Ft. Collins, Colorado. Interest in the use of contraceptives for controlling pest vertebrate species dates back to the 1960s (Balser, 1964). The public generally regards efforts toward reducing reproduction in overabundant vertebrates as a most acceptable approach. The terms contraceptives, chemosterilants, antifertility agents, and reproductive inhibitors are often used interchangeably when referring to the chemicals used to inhibit reproduction in vertebrate wildlife. There are many complex problems associated with the development of effective wildlife contraceptives and in making the approach practical for field use. Promising methods presently depend mostly on endocrine disruption and immunocontraceptives. Research has shown that to be effective and useful, a contraceptive should have the following characteristics: Be safe for the target animal Be free of undesirable side effects l Not cause treated food animals to be unsafe for human consumption l Cause little or no social effect to target animals l Induce complete and long-lasting infertility that, ideally, is reversible

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Only one contraceptive product is currently U.S. EPA registered for vertebrate pest control and that is OvoControl for wild Canada geese. It is intended for use on resident goose populations that build up in urban parks and other areas where they may present a nuisance or public health problem. This same material is being evaluated as a contraceptive bait for domestic pigeons in urban areas. Significant advancements in wildlife contraceptives are anticipated in the future.

6.10.5  Glue Boards Glue boards are designed to capture rats and house mice and consist of a 3/16- to 5/16-in. layer of sticky nonhardening adhesive with extraordinary holding properties. This adhesive is applied to a rectangular-shaped piece of cardboard or a similar tough material or placed in shallow plastic trays. The size of the glue board varies but generally ranges in the area of 2.5  6 in. for mouse-sized boards and 3.5  10 in. for rat glue boards. The chemical formulas used to prepare these essentially nontoxic glues (i.e., adhesives) vary and remain trade secrets of the manufacturers. Glue boards are considered a trap and therefore do not come under the U.S. EPA’s pesticide registration process. For use, glue boards are placed along walls or in areas where rodents are known to travel. When the rats or mice run over the glue boards, they are entrapped. In the struggle to escape, the rodents usually get their muzzles in the glue and suffocate. The glue boards and their contents are disposed of following use; they are not intended to be reused. Glue boards are used almost entirely indoors or under some kind of cover because they become less effective if they get damp or wet. Temperature extremes and an accumulation of dust on the surface reduce effectiveness. Special glue board covers are marketed that permit rodent entry and exclude pets. Covers also keep the entrapped rodents out of sight until disposal. Glue boards offer a nontoxic means of rodent control and can be very effective, especially for trapping house mice. They are used by the public as well as pest management professionals. Their use has increased greatly over the past 20 years and continues to grow. One drawback to using glue boards is that sometimes the trapped rodent does not die. In these cases, the trapped animal should be euthanized in as humane manner as possible such as CO2 euthanasia or quick cervical dislocation.

l l

6.10.6  Livestock Protection Collars The livestock protection (LP) collar, more commonly referred to as a toxic collar, is a relatively new device used to selectively kill livestock-depredating coyotes. Each collar, constructed to surround the neck of a goat or lamb, has

Chapter | 6  Vertebrate Pest Control Chemicals and Their Use in Urban and Rural Environments

within it two sealed pouches that together contain a small quantity (300 mg) of the “Restricted Use Pesticide,” compound 1080 (sodium fluoroacetate), formulated in a liquid carrier. Collars are placed around the necks of a target group of young goats or lambs. Coyotes usually attack the throat of their prey; in doing so, they puncture the collar and ingest a sufficient amount of the toxicant to be lethal. The coyote may not exhibit symptoms and die for several hours and therefore may have fed on its kill and left the scene before it succumbs. Toxic collars are highly specific, targeting only those coyotes actually preying on livestock. Unfortunately, toxic collars have many limitations as to where they can be most effectively used, in addition to the fact that they are expensive and that a goat or lamb must be sacrificed for every coyote taken. Only a few states have been authorized by the U.S. EPA to use toxic collars, and then only in very rural areas. Those individuals utilizing them must go through an approved training program and be specially certified to use this device. The use restrictions and recordkeeping requirements have made use of this selective tool a last resort only, where the depredating coyote cannot be controlled by the use of traps, snares, or other standard methods.

6.10.7  Toxicant Ejector Device The M-44 is a spring-activated device used to propel an orally active toxicant into the mouth of a coyote when the device is triggered by the biting and pulling behavior of the targeted animal. The relatively small stakelike device consists of a capsule holder, a spring-activated ejector mechanism, a capsule containing 0.9 g of powdered sodium cyanide mixture, and a 5- to 7-in.-long specially designed hollow stake into which the ejector mechanism is inserted. Sodium cyanide is a “Restricted Use Pesticide.” The M-44 is positioned and set just off the trails showing evidence of use by coyotes when entering the livestock area. In addition to coyotes, the device and sodium cyanide are also registered for taking red and gray fox. To set the device, the hollow stake is first driven into the ground. The trigger ejection mechanism is cocked, inserted, and secured inside the stake. The capsule holder, which has been wrapped with an absorbent material and loaded with a cyanide capsule, is screwed onto the positioned below-ground ejector unit. When set, only a few inches of the device projects above the ground surface. A small amount of fetid meat bait is applied to the absorbent wrapping surrounding the capsule holder. In addition to the baited device, a dab of coyote lure may be placed on a nearby bush to draw coyotes from a greater distance. Coyotes attracted by the bait will try to bite the baited capsule holder. In the process, they will pull on the exposed capsule holder and trigger the device. The spring-activated plunger forcefully propels the sodium cyanide from the capsule through the open end of the holder and into the coyote’s mouth. Death results within a few seconds.

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The M-44 is very selective for canids because of the baits and species-specific lures used and because the device is designed so that it can only be triggered by an upward pull. The device can be used with relative safety in pastures where livestock are present. Where M-44s are employed, the property is posted with warning signs to alert individuals to their presence. A special training program is required before the M-44s can be used. In some states, only federal employees involved in predator control are permitted to use the devices. In certain other states, the M-44s can be used by trained and certified livestock producers. The U.S. EPA has authorized the use of M-44s only in certain states that have a demonstrated need and have developed an appropriate training program.

6.10.8  Flock Dispersal Agent Avitrol (4-aminopyridine) is registered as a flockfrightening repellent and is used in a bait form to frighten pest bird species such as pigeons, house sparrows, and certain blackbirds and cowbirds from structures and the vicinity of structures. In agricultural situations, the Avitrol bait may be used for a somewhat broader group of birds. The material is formulated on grain baits. This treated grain is then blended with untreated grain to give the appropriate dilution. The dilution ratio may vary depending on the pest species. Such diluted baits are placed in trays or on rooftops accessible to the target pest species. A period of prebaiting with a placebo precedes exposure of the treated bait. The ingestion of an active amount of Avitrol by a small proportion of birds causes the affected birds to emit distress calls and display erratic behaviors that frighten away the remaining birds of the flock. The use of diluted baits limits the number of birds affected. The material is sufficiently toxic that some of the birds that are affected will succumb. Dead birds are immediately picked up following treatment. When the flocks are adequately frightened, the birds may not return to that area for months. Avitrol is a “Restricted Use Pesticide” and can be used by or under supervision of government agencies, by licensed and certified structural pest management professionals, and by certified applicators. It is not available to the public. Its use is relatively limited because a few birds may die outside of the treated property and this often results in an adverse public reaction. The use of the material is prohibited in some cities.

6.10.9  Poison Bird Bait Starlicide (3-chloro-p-toluidine hydrochloride) is registered and marketed for starling and blackbird control in and around livestock- and poultry-raising facilities. It is

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formulated as a grain-based pellet and used by or under the direction of personnel trained in bird damage control. The active ingredient is frequently referred to as DRC-1339 or Compound DRC-1339. USDA–APHIS presently maintains the registration of the DRC-1339 concentrate, and its use is restricted to APHIS personnel. The USDA–APHIS registration has been expanded to include several other pest species, namely, pigeons, gulls, ravens, crows, and magpies. In livestock feedlots and poultry operations, the bait is placed in feeding stations. Applications are made before the starlings and blackbirds arrive for their first morning feeding. Starlicide is ineffective for house sparrows and several other pest birds, as there is a wide variation in sensitivity to the toxicant among bird species. Hawks and mammals are relatively resistant to the material. APHIS personnel have used this material to control crows, ravens, and magpies that prey on the eggs or young of federally designated threatened or endangered species, as well as those preying on newborn livestock. Baits for these purposes may be prepared with eggs or meat. The use of Starlicide and baits prepared with DRC1339 concentrates is fairly limited; most of it is used in rural situations.

Conclusion Vertebrate pests are a diverse group of animals and include amphibians, reptiles, birds, and mammals. Most wildlife species are considered pests only in relatively few circumstances and may be neutral or highly desirable in most situations. The majority of vertebrate pest problems are prevented or resolved without the use of pesticides and/or lethal controls. A variety of approaches are employed, including exclusion by fencing, netting, or rodent and bird proofing of buildings; habitat modification; selective crop cultural practices that include resistant crop varieties; and pest-frightening methods. Several nonpesticide approaches such as hunting and shooting are employed to kill offending

animals or to reduce local pest populations. Vertebrate pesticides are represented by two major groups of agents, the lethal or toxic compounds selected or developed for killing the pest animal and the chemical repellents that are used to discourage pests from feeding or gnawing on some item or to repel them from a particular area. Specific pesticide approaches are used for rodents, birds, and predators. Novel approaches such as flock dispersing agents, contraceptives, and immobilizing are also used to solve specific vertebrate pest problems.

References Balser, D. S. (1964). Management of predator populations with antifertility agents. J. Wildl. Manag. 28, 352–358. Conover, M. (2002). “Resolving Human–Wildlife Conflicts.” Lewis Publishers, Boca Raton, FL. Corrigan, R. M. (2001). “Rodent Control: A Practical Guide for Pest Management Professionals.” GIE Media, Cleveland, OH. Hygnstrom, S. E., Timm, R. M., and Larson, G. E. (eds.) (1994). “Prevention and Control of Wildlife Damage.” Nebraska Cooperative Extension Service, University of Nebraska, LN, USDA–APHIS– Animal Damage Control, and Great Plains Agriculture Council. Marsh, R. E. (1986). Vertebrate pest management. In “Advances in Urban Pest Management” (G. W. Bennett and J. M. Owens, eds.), pp. 253–285. Van Nostrand-Reinhold, New York. Mason, J. R. (1998). Mammal repellents: options and considerations for development. In “Proceedings of the 18th Vertebrate Pest Control Conference” (R. O. Baker and A. C. Crabb, eds.), pp. 325–329. University of California, Davis, CA. National Academy of Sciences (1970). “Vertebrate Pests: Problems and Control.” National Academy of Sciences, Washington, DC. Salmon, T. P., and Dochtermann, E. A. (2006). Rodenticide grain bait ingredient acceptance by Norway rats (Rattus norvegicus), California ground squirrels (Spermophilus beecheyi) and pocket gophers (Thomomys bottae). Pest Manage. Sci. 62, 678–683. Salmon, T. P., Whisson, D. A., and Marsh, R. E. (2006). “Wildlife Pest Control Around Gardens and Homes,” Publication 21385, 2nd ed. University of California, Division of Agriculture and Natural Resources.

Chapter 7

Pesticide Use and Associated Morbidity and Mortality in Veterinary Medicine Robert H. Poppenga1 and Frederick W. Oehme2 1

University of California, Davis, California Kansas State University, Manhattan, Kansas

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7.1  Introduction Pesticides are frequently topically applied or orally administered to animals to control harmful insects and parasites or used in their environment to control a variety of pests. The rural setting of food-producing and livestock-rearing operations results in exposure of domestic animals to the wide array of agricultural chemicals currently in use. In addition, wildlife species are often exposed accidentally or maliciously to pesticides, especially those used in animal and plant agriculture. Exposure of wildlife can occur either directly or via their prey species. Pesticide exposures can be minimal or can be sufficiently great to produce clinical signs and result in acute poisoning, delayed toxicity, or residues that affect public safety through contamination of the food chain. Various publications have detailed the clinical problems resulting from exposure of domestic animals to pesticides (Antoniou et al., 1997; Oehme and Rumbeiha, 1999; Osweiler et al., 1985; Postgraduate Committee in Veterinary Science, 1987). Pesticide use can also indirectly affect wildlife as a result of elimination of pest species that might be important prey items. Two other aspects of pesticide use specific to veterinary medicine bear mentioning. First, the use of pesticides, especially herbicide use on lawns and insecticides used in flea and tick products, has been linked to the occurrence of certain cancers in domestic dogs (Glickman et al., 2004; Hayes et al., 1991; Raghavan et al., 2004). Although results from these studies suffer from the limitations common to many retrospective epidemiologic investigations (Kaneene and Miller, 1999), there is increased awareness on the part of many pet owners about possible adverse health effects to their pets from pesticide exposure. Second, there is concern about the adverse environmental impact of the routine

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veterinary use of some pesticides for the control of animal pests. A large number of new pesticides have replaced many older pesticides that previously had been used widely on animals or in their environments. Specifically, organophosphorus and carbamate insecticides are used much less frequently in veterinary medicine now than in the past as a result of regulations restricting their use. In their place, generally less toxic insecticides, such as pyrethrins and pyrethroids, neonicotinoids, and insect growth regulators, have been developed. Newer pesticides have been specifically designed for rapid killing of pests and relative safety to pets and people. The efficacy and safety of newer pesticides are a result of exploiting unique physiologic differences between mammals and specific pests such as insects, intestinal parasites, or ticks. This has resulted in fewer acute domestic animal intoxications occurring now than in the past, although acute intoxications still occur with regularity. Acute wildlife intoxications still commonly occur as a result of accidental exposure to a variety of pesticides, especially more toxic agriculturally used insecticides and rodenticides. Also, unfortunately, malicious poisoning of dogs, cats, and wildlife by a variety of pesticides is still common. Interestingly, animals can be exposed to a variety of naturally occurring chemicals with demonstrated or purported pesticidal activity. Although poorly documented, undoubtedly such exposures have become more common as people have become increasingly concerned about exposure to synthetic pesticides and due to the belief that naturally occurring pesticides are less toxic. Investigation into the repellent and insecticidal activities of a variety of plant-derived chemicals for use in veterinary medicine is ongoing (George et al., 2008).

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Although the impact of pesticides on wildlife is discussed more fully elsewhere, veterinary medicine is increasingly concerned about protection of our wildlife resources. Wildlife species intoxicated by various pesticides are routinely presented to wildlife treatment and rehabilitation centers. In addition, investigations of wildlife mortality events, which are frequently a result of pesticide exposure, are commonly conducted by veterinarians and veterinary diagnostic laboratories. This chapter focuses on the use of pesticides in domestic food and pet species.

7.2  Formulations There are a variety of pesticidal formulations designed to be administered to animals dermally, orally, or via injection. The formulations used depend on the animal species involved, label restrictions, the conditions of use, and other factors that determine the particularities of the desired effect and the most effective application procedure (Oehme, 1987c). Pesticides have been formulated for inclusion into sprays, dips, powders, collars, ear tags, feeds, pastes, gels, and spot-ons, among others. The introduction of injectable and pour-on avermectin insecticides has provided alternatives to older, conventional methods of insect control and allows for more sustained insecticidal action and less need for animal handling. The latter is an important consideration in situations in which it is difficult to handle animals on a regular basis, such as beef cattle kept on large, open grassland ranges. Of equal significance is the use of insecticide impregnated ear tags that contain the synthetic compounds permethrin and fenvalerate, which initially were introduced to control horn flies in range cattle. Dermal application of systemic organophosphorus (OP) compounds may also control parasites in remote anatomical locations, such as migrating stages of the cattle grub Hypoderma larvae (Arther and Shmidl, 1999). Market data from various animal health companies show wide acceptance of these products. Most of the OP and permethrin compounds are rapidly and completely absorbed from the source and distribute diffusely through the body systems, giving broad-spectrum insecticidal effects. However, unlike the organochlorines, these compounds are biologically unstable and do not accumulate in the environment. They have short half-lives and are eliminated relatively quickly from the body. Systemic OP compounds present some special problems if they are not applied at appropriate seasonal times, if they are used in overdose, or if they are applied to animals specifically sensitive to their properties (Osweiler et al., 1985). Systemic OPs must be applied during the fall season, when internal parasites are in their most vulnerable migratory patterns. If applied too early, the agents are less effective; if applied too late, the location of the parasite when attacked by the insecticide may produce serious and

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often fatal complications due to the release of toxins from dead larvae. Time of application obviously is critical when these compounds are used. Because of the rapid absorption and distribution of these compounds, dosage is critical. Overestimation of body weight or too zealous application of the formulation can produce acute toxicity.

7.3  Species sensitivities Pesticide use in domestic animals involves exposure of large numbers of different species, on each of which adverse effects can occur. In contrast to the single human organism, domestic animals have anatomical, physiological, and biochemical differences that can significantly alter the ultimate kinetics and, therefore, the efficacy and toxicity of a pesticide (Oehme, 1987a). Although anatomical and physiological species differences are most obvious, they are less significant than biochemical differences in digestive tract enzymes, circulating enzymes, and liver enzymes and variations in other degradative processes that affect the detoxification mechanisms that help animals deal more or less effectively with pesticides. Biochemical differences between breeds of the same species are expressed clinically as variations in sensitivity to pesticide exposure (Oehme, 1987a). Whereas monogastric animals (horse, swine, dog, and cat) have stomachs physiologically and biochemically similar to that of humans, cattle, sheep, and goats have a unique part of the digestive tract (the rumen) that serves as a fermentation vat for converting cellulose forage into protein precursors. Pesticides ingested by cattle, sheep, or goats immediately arrive in this large fermenting and reducing environment with a pH that may significantly modify the diffusability of the chemical. The pesticide is diluted in the 40–60 gallons of ruminant stomach content, where reductive biochemical action on the pesticide is rapid. In most cases, the result is the beginning of detoxification, although in certain instances activation can occur. Absorption through the rumen wall will be retarded until the pesticide or its metabolite comes into contact with the mucosa wall. Insecticides can be held in the fluid portion of the rumen material for many hours until rumen activity moves the chemical against the mucosa or into another portion of the digestive tract (Oehme and Barrett, 1986). Except in rare instances when highly irritating pesticides such as arsenical herbicides are ingested, ruminant animals do not vomit—a further difference from the normal physiological process expected in monogastric animals following exposure to such compounds. It is important to note that it takes several weeks for the digestive system of ruminants to develop after birth, with approximate adult distribution of stomach proportions being reached by approximately 8 weeks of age (Cunningham, 1997). Thus, young ruminants are more like monogastric animals than true ruminants,

Chapter | 7  Pesticide Use and Associated Morbidity and Mortality in Veterinary Medicine

and this can result in differences in sensitivity to xenobiotics following oral exposure. Several physiologic differences of birds increase their sensitivity to inhaled toxicants, such as a higher mass specific minute ventilation, a higher mass specific ventilation of gas-exchange tissues, cross-current and countercurrent gas exchange mechanisms, and a gas diffusion barrier onehalf the thickness of that of mammals (Brown et al., 1997). Some of these physiologic adaptations are as a result of high metabolic rates of birds and the concomitant need for a high ventilatory capacity. Of all the factors that influence the kinetics of xenobiotics (e.g., drugs or pesticides) among animals, differences in the rate of elimination, particularly for chemicals that undergo extensive hepatic biotransformation, generally account for species variations in xenobiotic disposition (Baggot, 2001). The biotransforming enzyme activities of ruminants are especially high in sulfatase activity and the ability to form sulfate conjugates, whereas pigs are especially active in glucuronidase activity and the ability to produce conjugates of glucuronic acid. Horses and dogs tend to have comparable oxidative mechanisms, leading to good sulfate and glucuronide formation, and are capable of biotransforming pesticides normally detoxified and excreted by those biochemical pathways (Oehme, 1987a). However, dogs are not able to acetylate and cats are relatively deficient in their ability to glucuronidate (Baggot, 2001). Even within the same species, breed differences result in differing sensitivities to xenobiotics such as pesticides. For example, it is well-known that Collie dogs and related breeds, such as Australian shepherds, possess a mutation in the canine multidrug resistance gene, MDR1, which encodes p-glycoprotein, an ATP-dependent drug transporter that moves a broad spectrum of substrates across several important tissue borders (Neff et al., 2004). A lack of p-glycoprotein at the blood–brain barrier in dogs possessing the mutation results in higher brain concentrations of some xenobiotics. The sensitivity of Collies and related breeds to ivermectin is due to the lack of p-glycoprotein (Hopper et al., 2002; Neff et al., 2004). Although there are well-known differences in sensitivity to xenobiotics due to sex within many species, such differences have not been well studied in livestock and pet species. However, one example is that of the OP chlorpyrifos, which is more toxic to adult bulls than to young bulls, cows, or steers (Haas et al., 1983). Higher testosterone concentrations were associated with a greater depression of cholinesterase activity, most likely due to an increased rate of production or decreased rate of metabolism of the oxon metabolite. With some notable exceptions, there is a general trend of shorter xenobiotic half-lives in herbivorous species such as cattle and horses than in carnivorous species such as dogs and cats (Baggot, 2001). Xenobiotic half-lives tend to be longer in humans than in domestic animals.

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Another factor that can influence the likelihood of intoxication is variation in behavior between species. Unlike dogs, cats are less likely to be poisoned by pesticides by virtue of their discriminating eating habits. However, they endlessly groom themselves, and any chemical that contacts their fur or their feet is carefully licked and swallowed. Adequate comprehension of the variability in toxicity from pesticides in domesticated animal species requires understanding the anatomy, physiology, and biochemistry of the exposed animals. Other factors that can affect the outcome of any toxicity include age, sex, health, nutritional status, and synergistic or antagonistic actions of other concurrently exposed chemicals. These factors often play vital roles in individual species sensitivities and the ultimate outcome of the therapy and management of toxicosis.

7.4  Pesticide use in domestic animals Use of pesticides in food-producing animals is economically necessary. Pesticides are required to control insects and parasites that reduce productivity in livestock, eat crops that are used for livestock feed, and carry diseases. In small animals, the use of pesticides is often esthetic, to keep dogs and cats from carrying unsightly and annoying fleas and ticks, but pesticides are also used to control internal parasites. The actual chemical(s) employed will vary with the specific circumstances of need. Current information on the quantities of pesticides used in the context of pest control in or on animals is difficult to find. It has been estimated that approximately 194 tons of parasiticides are used each year in Europe, but data specific to an individual active ingredient are limited (Boxall et al., 2009). In sheep, diazinon has been widely used, whereas in cattle the most widely used parasiticide is ivermectin, followed by oxfendazole, eprinomectin, doramectin, and fenbendazole (Boxall et al., 2007). Morantel, moxidectin, and permethrin are used in much lower amounts.

7.5  Regulation of pesticides used   in veterinary medicine Each country has its own unique regulatory scheme pertaining to pesticides used in veterinary medicine. In the United States, three federal agencies are involved in pesticide regulation: the Environmental Protection Agency (U.S. EPA), the Food and Drug Administration (FDA), and the Department of Agriculture (USDA). Cooperation among the three agencies is addressed by a memorandum of understanding (see http://www.fda.gov/AboutFDA/Partn ershipsCollaborations/Memorandaof UnderstandingMOUs/ DomesticMOUs/ucm116368.htm). In addition to federal regulatory oversight, each state has its own statutes and

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regulations concerning pesticides. Discussion of specific federal and state regulatory details is beyond the scope of this chapter. The U.S. EPA is responsible for administering and enforcing the Federal Insecticide, Fungicide, and Rodenti­ cide Act (FIFRA). Under this act, the U.S. EPA has the authority to protect humans and their environment from unreasonable adverse effects of pesticide chemicals by regulating the sale and use of pesticide products. Every U.S. EPA-registered pesticide product has a U.S. EPA registration number on its packaging. The U.S. EPA samples chemicals to verify label claims concerning content and safety, and it investigates incidents in which the misuse of pesticides may have occurred. The U.S. EPA is responsible under the Federal Food, Drug, and Cosmetic Act (FFDCA) for establishing tolerances and recommending action levels to the Food Safety Inspection Service (FSIS) of the USDA and FDA for residues of pesticides in food and has the authority to monitor the effectiveness of surveillance and enforcement. If evidence arises to challenge the safety of a registered pesticide product, the U.S. EPA reviews scientific data and takes action if necessary to reduce or eliminate the risks. Some veterinary products containing pesticides are considered to be drugs that are regulated by the Center for Veterinary Medicine (CVM) within the FDA. Another exception to pesticide registration requirements pertains to pesticides that the Administrator, under FIFRA, has determined “to be of a character which is unnecessary to be subject to this Act,” and that have been exempted from the requirements of FIFRA by regulation. In 1996, the U.S. EPA exempted certain minimumrisk pesticides such as eugenol, garlic, and mint oil from FIFRA requirements if they satisfy certain conditions. Many products used on animals or in an animal’s environment are available in stores or via the Internet. The U.S. EPA exempted such products in part to reduce the cost and regulatory burdens on businesses as well as the public for pesticides posing little or no risk and also to focus the U.S. EPA’s limited resources on pesticides that pose greater risk to humans and the environment. It is important to note that even if a pesticide product meets the conditions for exemption from regulation under FIFRA, it is still subject to any applicable requirements of the FFDCA if its use results in pesticide chemical residues on or in food commodities or animal feed. In addition, producers of pesticides must meet any applicable state registration or other regulatory requirements (see http://www.epa.gov/opp00001/about). The U.S. FDA is charged with the enforcement of the FFDCA. Under this act, the FDA is responsible for ensuring that human foods and animal feeds are safe and, among other things, that they do not contain illegal residues of drugs, pesticides, or environmental contaminants. The FFDCA authorizes the U.S. EPA to establish a tolerance for the maximum amount of a pesticide residue that may be legally present in or on a raw agricultural commodity.

Hayes’ Handbook of Pesticide Toxicology

This section also authorizes the U.S. EPA to exempt a pesticide residue in a raw agricultural commodity from the requirement of a tolerance. A tolerance or tolerance exemption is required when the U.S. EPA grants registration under FIFRA for the use of a pesticide in food and feed production in the United States. Registration of a pesticide is not, however, a prerequisite for establishing a tolerance. For example, the U.S. EPA may establish a temporary tolerance to permit the experimental use of a nonregistered pesticide, or it may establish a tolerance for a pesticide residue resulting from the use of the pesticide in food or feed production originating in a foreign country. Tolerances and exemptions from tolerances established by the U.S. EPA for pesticide residues in raw agricultural commodities are listed in 40 CFR Part 180. The FDA is responsible for the enforcement of pesticide tolerances and food additive regulations established by the U.S. EPA. This enforcement authority is derived from section 402(a)(2)(B) of the FFDCA. Under this section, a raw agricultural commodity or a processed food or feed is deemed to be adulterated and subject to FDA enforcement action if it contains either (1) a pesticide residue at a level greater than that specified by a tolerance or food additive regulation or (2) a pesticide residue for which there is no tolerance, tolerance exemption, or food additive regulation. Chemicals used to control diseases of humans or animals (e.g., livestock and pets) are not considered pesticides; such chemicals are regulated by the FDA as drugs. The USDA is charged with the enforcement of the Federal Meat Inspection Act (FMIA), the Poultry Products Inspection Act (PPIA), and the Egg Products Inspection Act (EPIA). Within the USDA, the FSIS is responsible for the wholesomeness and safety of meat, poultry, and products thereof intended for human consumption. This is accomplished, in part, by inspection at slaughtering and processing establishments and by sampling and analyzing edible tissues derived from livestock and poultry at the time of slaughter or after slaughter at other locations outside the establishment to ensure, among other things, that meat and poultry do not contain residues of drugs, pesticides, or environmental contaminants that cause them to be adulterated under FMIA or PPIA. The Agricultural Marketing Service is responsible for the wholesomeness and safety of egg products. It conducts inspection and samples for such residues at plants processing egg products to ensure compliance with EPIA. Because of safety concerns, some pesticides are not available to the general public in the United States. The “Restricted Use” classification restricts a product, or its uses, to use by a certificated pesticide applicator or under the direct supervision of a certified applicator. Currently, veterinarians are exempt from certification in their usage of restricted-use pesticides. In addition, under the Animal Medicinal Drug Use Clarification Act of 1994, licensed veterinarians can use and prescribe, under specified

Chapter | 7  Pesticide Use and Associated Morbidity and Mortality in Veterinary Medicine

conditions, animal and human drugs for extra-label use (Sundloff, 2009). Thus, pesticides regulated as drugs by the FDA can be used by licensed veterinarians in an extralabel manner. An example of extra-label drug use involves the oral or dermal use of certain macrolide endectocides in goats at higher than label dose rates due to their low systemic availability in that species compared to cattle and sheep (Lanusse et al., 2009). Such extra-label use of a drug requires caution on the part of the veterinarian to ensure that no violative drug residues occur. This generally involves determination of an extended withdrawal time. The Food Animal Residue Avoidance Databank (FARAD) is designed to provide veterinarians with guidance in establishing such withdrawal times (Riviere and Sundloff, 2009). Mechanisms for reporting suspected adverse reactions in animals to pesticides have been established. Veterinarians can report pesticide-related incidents involving domestic pets through a reporting page developed by the National Pesticide Information Center and the U.S. EPA (see http://pi.ace.orst.edu/vetrep/). The CVM also has a mechanism for reporting adverse drug events (see http://www.fda.gov/AnimalVeterinary/SafetyHealth/ ReportaProblem/ucm055305.htm) (Post, 2009). Adverse events can be reported to a product’s manufacturer as well; in turn, manufacturers are required to submit reports of adverse events to the U.S. EPA. Appropriate reporting of adverse events has resulted in increased scrutiny of the use of flea and tick control products on dogs and cats (see http://www.epa.gov/opp00001/health/flea-tick-control. html). More than 44,000 potential incidents associated with registered spot-on products were reported to the U.S. EPA in 2008.

7.6  Violative residues The ultimate challenge in using pesticides in domestic animals is to avoid the occurrence of chemical residues in animal products intended for human consumption. To this end, storage and excretion are studied in domestic animals early in pesticide development and are studied in the field as each pesticide is used in agricultural production. These are traditional tests mandated by regulatory and public health concerns. If properly conducted, such studies provide guidelines and appropriate safety margins for the use of pesticides in food-producing animals so that pesticide residues do not occur in foods for human consumption. Of practical interest is the continuing monitoring of meat and dairy products intended for human consumption (Somogyi et al., 1978). These “food basket” studies ensure that pesticides are used in agricultural practice in accord with label recommendations and sound chemical application. The occasional misuse of such pesticides is usually quickly detected through such monitoring assays (Clark et al., 1974), and the regulatory action taken to confine the

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contamination and eliminate its access to human foods is a tribute to the diligence of the FSIS of the USDA. The high volume of uncontaminated quality foods available to the U.S. public also validates the agricultural and veterinary professions’ recognition of the need for quality foods and their cooperative efforts to use pesticides appropriately in food-producing animals.

7.7  Frequency of intoxication There is no comprehensive database to determine the frequency of pesticide intoxication in animals, especially livestock. Veterinary hospitals and institutions have numerous reports of animals being poisoned, some fatally, following pesticide exposure. However, this information has not been fully integrated with other animal data and animal diseases in private and government collections to complete the circle and provide total realistic pesticide profiles. Fortunately, some information is available that sheds light on the relative frequency of pesticide intoxication in domestic animals compared to other toxicant categories. For example, the Pet Poison Helpline reported that of approximately 9000 animal toxicology calls, nearly 1800 were related to exposure of pets to house, lawn, and garden products, most of which were pesticides (Hovda, 2009). Exposure to veterinary products, which include topical flea and tick products and anthelmintics, accounted for approximately 500 calls. Pyrethrins or pyrethroid expo­sures accounted for more than 80% of the calls specifically related to insecticides, followed by calls related to avermectins (primarily ivermectin), imidacloprid, and boric acid. Outcomes varied depending on the toxicity of the specific compound, the amount ingested, and the species exposed. Exposure to cholinesterase-inhibiting insecticides (carbamates and organophosphorus insecticides) and arsenic-containing products was more often associated with clinical signs. Although the number of calls involving cats was relatively small compared to the number involving dogs, cats were more likely to show severe symptoms. Seasonal differences were also noted in terms of frequency of calls: Call numbers were highest during summer months and lowest during winter months. The majority of rodenticide-related exposures involved long-acting anticoagulant rodenticides, although exposure to zinc phosphide also occurred. The overwhelming majority of calls pertaining to rodenticides involved dogs. Clinical signs were unlikely to occur following exposure to herbicides. In 1995, a regional human poison control center serving veterinarians and pet owners received 6091 veterinary-related calls, of which 20.1% involved the exposure of pets to pesticides (Hornfeldt and Murphy, 2000). Similar percentages have been reported in other countries. In Brazil, dogs were intoxicated more frequently than cats. Rodenticides, farm pesticides, and household pesticides accounted for 15.8, 13.9,

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and 5.0%, respectively, of the dog poisonings and 10.6, 27.6, and 14.9%, respectively, of the cat poisonings (Xavier and Kogika, 2002). In dogs, the most frequently implicated farm pesticides were the OPs, carbamates, and amitraz, whereas in cats the most frequently implicated were the OPs and carbamates. Household pesticide categories were not provided. The incidence of livestock and poultry poisonings in five European countries has been assessed (Guitart et al., 2009). Many calls involved horses and cows exposed to herbicide spray drift, although confirmation of intoxications was hampered by lack of species-specific toxicity data and analytical tests to confirm exposure. In 2003, pesticide poisoning represented 33 and 50% of the confirmed poisoning cases in food-producing animals as recor­ ded by the Laboratory of Toxicology at Ghent, Belgium, and the Centre National d’Informations Toxicologiques Vétérinaires in France, respectively (Guitart et al., 2009). Although historically exposure to lindane and endosulfan was relatively frequent, banning or restricting the use of these organochlorine insecticides has decreased the incidence of intoxication. Metaldehyde and methiocarb, used for slug and snail control, are associated with cattle and sheep poisonings (Royer and Buronfosse, 1998). Chlorate was the most commonly implicated herbicide in livestock poisonings. In studying fatal animal poisonings in northern Greece from 1990 to 1995, 926 animal tissues were analyzed by chromatographic techniques. Pesticides caused 78% of the poisoning cases, whereas all other toxic substances caused 22%. The animals affected were mainly cats, dogs, sheep, birds, and bees (Antoniou et al., 1997). Carbamates (methomyl, carbofuran, carbaryl, aldicarb, and mecarbam) were responsible for 46.9 and 66.7% of the sheep and goat poisonings, respectively. The herbicides paraquat and atrazine were implicated in 7% of the small ruminant (sheep and goats) poisonings. Despite the lack of more specific statistics, it seems certain that pesticide poisoning is responsible for a higher proportion of unintentional mortality in domestic animals than in humans. This is true because animals have a greater exposure to pesticides. Animals will be injured as a result of intentional application of pesticides at a greater rate than humans because treatment of animals with pesticides is far more common and aggressive, and there can be no real cooperation on the part of the animals to avoid errors in pesticide mixing or application. Treating uncooperative larger animals often necessitates using such methods as large-scale spraying or dipping, which present physical as well as chemical dangers to the animals. Specific dosages and application rates are often generalized, and weak or highly susceptible individuals in a group run high risks. In Spain, the frequency of pesticide intoxication of domestic animals was inversely related to the pesticide LD50s but not the relative amounts of specific pesticides used

Hayes’ Handbook of Pesticide Toxicology

(Martínez-Haro et al., 2008). Also, malicious use of pesticides was not related to whether their use was restricted. Weighted against the disadvantage that toxic effects can occur through the use of pesticides is the fact that insects, parasites, and fungi are more important sources of agricultural and livestock losses and cause economic loss and diseases in domestic animals and livestock.

7.8  Scenarios of concern 7.8.1  Acute Intoxication In the vast majority of poisonings, the circumstances of exposure of domestic animals and the inherent toxicity of pesticides result in acute toxic effects. Clinical signs often appear within hours after exposure and dramatically demonstrate the biological effectiveness of these poisons. Symptoms can vary from the neurological signs seen with most of the insecticides to the bleeding and hematological abnormality seen with anticoagulant rodenticide poisoning. With rare exceptions, the attending veterinarian finds several animals already dead and others showing clinical signs of massive overexposure. The effects are similar to those expected in humans for the specific insecticide, fungicide, or rodenticide involved (Osweiler et al., 1985). Acute poisoning by pesticides is rare when they are used in accordance with manufacturer instructions and governmental regulations. This is because careful evaluation of pesticides occurs before they are licensed for use, and pesticides should have moderate safety factors in nontarget species if they are to be used as applications on animals. This has resulted in a surprisingly high degree of safety in their intended use. During a 4-year period when dipping more than 1 million cattle per preparation was performed by officials of the Tick Quarantine Area of New South Wales, the mortality due to poisoning was 0.0061% for coumaphos, 0.00342% for dioxathion, and 0.0349% for ethion. The overall mortality rate among more than 17.5 million head of cattle dipped in that process was 0.00322%—a percentage that would be unacceptable for humans but is surprisingly low given the agricultural circumstances (Roe, 1969).

7.8.2  Chronic Intoxication Chronic pesticide intoxication, defined by long-term lowlevel exposure, is relatively rare in animals compared to acute intoxication. However, prolonged recovery following acute intoxication is not uncommon following intoxication with some pesticides. For example, after resolving the acute signs of organophosphate poisoning, a further distinct manifestation of exposure to these compounds can occur in humans (Senanayake, 1998; Senanayake and Karalliedde, 1987) and in animals. The syndrome consists

Chapter | 7  Pesticide Use and Associated Morbidity and Mortality in Veterinary Medicine

of neurological signs that develop 24–96 h after acute exposure and produce a proximal paralysis, the progression of which is not altered by atropine or oxime treatment. An additional third syndrome produced by some organophosphate compounds is organophosphate-induced delayed neurotoxicity (OPIDN), which is a symmetrical sensory–motor axonopathy that tends to be most severe in the long axons and occurs 7–14 days after insecticide exposure (Davis et al., 1999). The initial event in the pathogenesis of this syndrome appears to be inhibition of neuropathy target esterase. This condition has been described in cattle in which nonpesticide phosphate esters induced delayed neurotoxicity (Coppock et al., 1995). A case of presumptive OPIDN in a bull has also been described (Perdrizet et al., 1985). The signs included recumbency with severe symmetrical paresis of all four limbs, where the hindlimb involvement was greater than that of the forelimbs. Cerebrospinal fluid analysis was normal. The bull was killed and necropsied. There were no gross central nervous system lesions. Histopathologic findings were similar to those in delayed neurotoxicity caused by organophosphate compounds. The bull had been treated with an organophosphorus insecticide, famphur, 43 days prior to the onset of signs. In addition to OPIDN, there is an intermediate syndrome that is most commonly seen with more lipophilic OPs and has been most commonly noted in cats (Mensching and Volmer, 2007). The intermediate syndrome is believed to result from downregulation of muscarinic receptors with sublethal, prolonged exposures. Clinical signs are primarily a result of nicotinic receptor stimulation because nicotinic receptors are not downregulated. Signs occur within 3–10 days of exposure and include weakness—in part manifested as neck ventroflexion due to the lack of a nuchal ligament in cats—anorexia, muscle tremors, depression, and death (Blodgett, 2006).

7.8.3  Pesticide Use and Cancer in Animals The association between pesticide exposure and the occurrence of various human cancers has long been debated and studied. Studies linking the occurrence of malignant lymphoma in dogs with the owner’s use of 2,4-D and of seminomas in military working dogs in Vietnam to pesticide exposure raised the issue of an association in veterinary medicine (Hayes et al., 1990, 1991). A review of the malignant lymphoma study refuted an association with the use of 2,4-D because of numerous limitations to the original study design (Carlo et al., 1992). An epidemiologic study of household dogs reported an increased risk of bladder cancer with topical insecticide use (Glickman et al., 1989). Studies have also associated an increase in transitional cell carcinomas in Scottish terriers with the use of topical spoton flea control products and exposure of lawns to herbicides (Glickman et al., 2004; Raghavan et al., 2004).

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Unfortunately, there are not enough data to draw any firm conclusions concerning pesticide exposure and cancer (or other disease processes) in animals. However, such studies do emphasize the potential importance of domestic animals as sentinels for human diseases. Animals could help monitor the environment, giving early warnings of environmental contamination. Information from pet-owning households can usefully be incorporated into research projects and even the census of the human population. Collecting information on the medical histories and behavior of pets does not require deliberate experimentation on animals. Scientists at the National Research Council (1991) recommended that government and other institutions develop pet animal population surveys and structure investigations of their diseases and exposure to toxins to monitor human and environmental health. Although people do not keep pets to monitor the environment, studying companion animals and livestock is invaluable for several reasons. Pet owners share living space with their animals: The pet dog or cat actually lives closer to the baby or toddler than to the adult human airspace. Second, animals may be more sensitive to specific pesticides and easily poisoned by conditions that seem safe to human beings. Third, animal diseases tend to progress at more rapid rates than the same disorders in humans, so syndromes can be identified and studied earlier and the results can be extrapolated quickly to humans. Finally, veterinarians see a majority of the pets in the United States (70%), and their medical histories would be readily available for evaluation.

7.8.4  Pesticide Use and Exposure of People in Contact with Animals An aspect of human pesticide exposure that has recently received attention involves exposure to pesticides via contact with animals. It has been estimated that 30% of all households own one or more pet dogs and that 50% of dog-owning households have at least one child living in them (American Veterinary Medical Association, 1997). This means that millions of children could be in contact with flea control insecticides via contact with their dogs. The potential for exposure of pet owners to the OP tetrachlorvinphos (TCVP) via use of the insecticide in flea control collars was investigated by Davis et al. (2008). TCVP residues were detected on glove and t-shirt samples, and a TCVP metabolite was detected in the urine of household members. The toxicologic significance of the exposure was uncertain, but the authors cautioned that direct contact by children to treated pets should be minimized for the first few days postapplication. Exposure of people via their pets to diazinon applied to lawns has also been studied (Morgan et al., 2008). The authors concluded that pet dogs appear to be an important pathway for the transfer of diazinon residues into homes.

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7.9  Major pesticide categories 7.9.1  Cholinesterase Inhibitors: Organophosphoruses and Carbamates Due to their more rapid breakdown in the environment, OP and carbamate insecticides replaced the organochlorine pesticides, which were banned for use in North America and Europe in the 1960s and 1970s. A number of different formulations are available for use either in the environment (e.g., agricultural or residential use) or on animals (e.g., livestock dips or sprays). OPs and carbamates are formulated as liquids, granules, and powders. The more toxic insecticides of each group are generally restricted to agricultural uses, whereas less toxic members are (or have been) approved for use on animals or in residential environments. Cholinesteraseinhibiting insecticides have been widely used either on animals or in their environment, and historically intoxications have been relatively common, especially in avian wildlife. For example, at the height of use of granular formulations of the carbamate insecticide carbofuran, 17–91 million birds were estimated to have died annually (Mineau, 2005). However, in recent years, their use, especially the use of OPs in and around homes, has declined due to concerns about human and wildlife exposure and resultant increased regulatory scrutiny and restrictions (Phillips, 2006). Possible exposure scenarios are numerous. Animals can be exposed via their diets, via home or premise use, or via direct application. Wildlife species are often exposed as a result of scavenging contaminated carcasses. Inhalation exposure is also possible from the use of dichlorvos-impregnated pest strips or premise spraying or fogging. Although chemically distinct, the OPs and carbamates have a common mechanism of toxic action, namely the inhibition of cholinesterase enzymes. Whereas acute intoxications are more common in animals, more chronic intoxications can occur, as previously mentioned. There are numerous cases of intoxication of wildlife documented in the literature but, interestingly, relatively few involving domestic animal species. Although published cases involving domestic animals are relatively few, the relative paucity belies the frequency of intoxication. For example, 162 cases of aldicarb intoxication were diagnosed during a 10-year period by the University of Georgia Veterinary Diagnostic and Investigational Laboratory (Frazier et al., 1999). Dogs were most often involved, but cats, horses, cattle, and goats were also poisoned. Most cases involved the intentional poisoning of animals via spiking of foods such as frankfurters, ham, and ground beef. The authors concluded that the cases presented to the diagnostic laboratory for investigation were only a fraction of those likely to have occurred. The following cases serve as specific examples of how domestic and wild animals have been poisoned. An accidental poisoning of 28 Holstein cows occurred in 1997 when 0.9 kg of 25% active ingredient fonofos (O-ethyl

Hayes’ Handbook of Pesticide Toxicology

S-phenyl ethylphosphonothiolothionate) was spilled onto bulk feed in a delivery truck. Eight cows died within 2 days; the remaining 20 were necropsied 29 days later. Of the 8 fatally poisoned, 7 were being fed a high-grain diet and 1 was fed a medium-grain diet. Fonofos concentrations in feed cart and storage bin samples were 100 and 61 g/g, respectively (Kurtz and Hutchinson, 1982). The oral acute LD50 and LD1 of fonofos in Holstein cows were calculated by the Litchfield and Wilcoxon method to be 1.30 and 0.84 mg/kg, respectively, with a 95% confidence interval of 0.20 mg/kg. Nine fairways of a golf course in Bellingham, Washington, were treated with diazinon AG500 at a target application rate of 2.2 kg active ingredient per hectare. Eighty-five American pigeons (Anas americana) died after grazing on one treated fairway on the day of application following irrigation (Kendall et al., 1992). The brains of all 85 pigeons were analyzed for acetylcholinesterase activity. Pigeons that died in the study area (n  85) had 44–87% depression of acetylcholinesterase activity (mean, 76%; SD, 7.1%) compared to control pigeons (n  3). Upper digestive tract contents of 15 of the 85 dead pigeons contained 0.96–18.1 ppm diazinon. American pigeons appear to be vulnerable to diazinon exposure. An unmeasured amount of 57% malathion added to a dipping solution that had been diluted by rain led to the poisoning of 24 dogs treated for ticks. Of these dogs, 16 died or had to be killed for humane reasons. The final concentration of the dip could have been as high as 15% malathion compared to the 0.25–0.5% concentration commonly used for this purpose (McCurnin, 1969). Two cats were exposed to chlorpyrifos used to overzealously spray an apartment for fleas (Jaggy and Oliver, 1990). Both cats developed clinical signs consistent with an intermediate syndrome as has been described in people. Both cats responded to treatment with pralidoxime and atropine. The likelihood of significant residues in milk, meat, or eggs from OP- or carbamate-exposed livestock is relatively low due to the rapid metabolism of the insecticides. A review of OP elimination from animals concluded that milk residues following OP treatment at recommended rates were low (Osweiler et al., 1985). In general, residues of OPs are less than 0.01 ppm 3–7 days after treatment. Specific times to reach such levels depend on variable factors, such as specific insecticide involved, the dose, and the route of exposure. The death of intoxicated animals can present problems associated with carcass disposal, and appropriate environmental precautions are necessary; consultation with local or state environmental agencies is recommended.

7.9.2  Pyrethrins and Pyrethroids There are a variety of pyrethrins and pyrethroids for use on animals or in their environment. They are marketed in

Chapter | 7  Pesticide Use and Associated Morbidity and Mortality in Veterinary Medicine

a variety of formulations, including sprays, dusts, dips, shampoos, spot-ons, gels, foggers, ear tags, pour-ons, and back and face rubbers (Volmer, 2004). Most ready-to-use formulations are at concentrations of 2% or less, although more concentrated formulations are available such as spoton permethrin products for dogs, which are available overthe-counter at 45–65% concentrations. Systemic absorption of dermally applied pyrethrins and pyrethroids is low (2%), which contributes to their low mammalian toxicity (Wollen et al., 1992). However, there can be inadvertent oral exposure via grooming or inhalational exposure. Fish are highly sensitive to pyrethrins and pyrethroids. Environmental contamination of bodies of water should be avoided. In homes, exposure of aquarium fish can occur when the premise is sprayed or fogged. This can occur in covered aquaria if the aerator is left on (Ensley, 2007). The tank and aerator should be covered during insecticide use and treated areas well ventilated prior to uncovering the tank and starting the aerator pump. Pyrethrins and pyrethroids are considered to be safe for use around avian species, although carriers and propellants in spray formulations can present inhalational hazards (Ensley, 2007). Although these insecticides should also be safe for use around exotic animals, species-specific toxicity data are lacking. Pyrethrins and pyrethroids are widely used on and around small companion animals such as dogs and cats. Ready-to-use formulations with active ingredient concentrations of 2% or less pose little hazard for use on dogs or cats. However, concentrated permethrin spot-on products are quite toxic for cats (Dymond and Swift, 2008; Meyer, 1999; Richardson, 2000). Between January 1994 and August 1998, the U.S. EPA received 95 reports of permethrin toxicosis involving 125 cats treated with concentrated spot-on products (Meyer, 1999). Thirty-three cats died or were euthanized. The products were applied by individuals who did not realize that concentrated permethrin was potentially toxic to cats. Misapplication was attributed to inadequate package warnings, confusion among spot-on flea control products with regard to active ingredients, and products being available with similar brand names but different active ingredients (Meyer, 1999). Interestingly, an additional 24 cats were believed to have been intoxicated as a result of secondary exposure to permethrin-treated dogs. Pyrethrin or pyrethroid intoxication of livestock is rare. Several studies have investigated the potential for residues of pyrethroids in milk of dairy animals under typical conditions of use. Depending on the specific insecticide evaluated, residues in milk can persist for several weeks after dermal application or use in impregnated ear tags (Bissacot and Vassilieff, 1999; Braun et al., 1985). The ecotoxicologic effect of fecal pyrethroid residues has been studied for some compounds. Depending on the time during the life cycle of the insect that exposure occurs

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and the number of treatments, the effect of deltramethrin on dung beetles in cattle feces ranged from negligible to almost complete elimination (Wardhaugh et al., 1998).

7.9.3  Natural Products Used   for Flea Control 7.9.3.1  d-Limonene and Linalool These insecticides are volatile oil extracts from orange peels. Insecticidal action is due, at least in part, to elaborated vapors (Baynes, 2009). Shampoos and sprays are available for use on dogs and cats. Although considered to be relatively safe, adverse effects can occur in cats, and some citrus oil formulations or use of pure citrus oil may pose a poisoning hazard (Lee et al., 2002; Powers et al., 1988). Fatal adverse reactions have been reported in cats following the use of an “organic” citrus oil dip (Hooser et al., 1986).

7.9.3.2  Melaleuca Oil Melaleuca oil is derived from the leaves of the Australia tea tree (Melaleuca alternifolia) and is often referred to as tea tree oil. The oil contains terpenes, sesquiterpenes, and hydrocarbons. A variety of commercially available products contain the oil, and shampoos and the pure oil have been sold for use on dogs, cats, ferrets, and horses. Tea tree oil toxicosis has been reported in dogs and cats (Bischoff and Guale, 1998; Villar et al., 1994). A case report describes the illness of three cats exposed dermally to pure melaleuca oil for flea control (Bischoff and Guale, 1998). Clinical signs in one or more of the cats included hypothermia, ataxia, dehydration, nervousness, trembling, and coma. Two cats recovered within 48 h following decontamination and supportive care. However, one cat died approximately 3 days following exposure. The primary constituent of the oil, terpinen-4-ol, was detected in the urine of the cats. Another case involved the dermal application of seven or eight drops of oil along the backs of two dogs as a flea repellant (Kaluzienski, 2000). Within approximately 12 h, one dog developed partial paralysis of the hindlimbs, ataxia, and depression. The other dog only displayed depression. Decontamination (bathing) and symptomatic and supportive care resulted in rapid recovery within 24 h.

7.9.3.3  Pennyroyal Oil Pennyroyal oil is a volatile oil derived from Mentha pulegium and Hedeoma pulegiodes. Pennyroyal oil has a long history of use as a flea repellant and has been used to induce menstruation and abortions in humans. There is one case report of pennyroyal oil toxicosis in the veterinary literature in which a dog was dermally exposed to pennyroyal

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oil at approximately 2 g/kg (Sudekum et al., 1992). Within 1 h of application, the dog became listless and within 2 h began vomiting. Thirty hours after exposure, the dog exhibited diarrhea, hemoptysis, and epistaxis. Soon thereafter, the dog developed seizures and died. Histopathologic examination of liver tissue showed massive hepatocellular necrosis. The toxin in pennyroyal oil is thought to be pulegone, which is bioactivated to a hepatotoxic metabolite called menthofuran.

7.9.4  Macrocyclic Lactones The avermectins and milbemycins (macrocyclic lactones or MLs) are closely related 16-member macrocyclic lactones produced through fermentation by soil-dwelling Streptomyces. The compounds used most commonly in veterinary medicine include abamectin, ivermectin, eprinomectin, doramectin and selamectin, and milbemycin oxime. These compounds are efficacious against a number of economically important nematodes and arthropods of animals as a result of their exceptional potency, high lipophilicity, and prolonged persistence of their activity. When used according to label directions, avermectins and milbemycins have a good safety record. Large safety margins are due to the selectivity of their pharmacodynamic action (Lanusse et al., 2009). MLs exert their parasiticidal effects via GABA agonism. GABA receptors are found in the peripheral nervous system of nematodes and arthropods, whereas in mammals they are found in the central nervous system. The relative inability of MLs to cross the blood–brain barrier of mammals due to the presence of p-glycoprotein accounts for comparative low mammalian toxicity compared to their parasiticidal activity. In ruminants, MLs are formulated for subcutaneous, oral, or topical administration (Lanusse et al., 2009). In swine, MLs are approved for use via subcutaneous, intramuscular, and oral feeding formulations. In horses, formulations are restricted to orally administered pastes or gels. In dogs and cats, chewable, topically applied, and injectable formulations are available. Not all MLs are approved for use in all species. For example, in horses only ivermectin and moxidectin are approved for use, whereas in dogs ivermectin, selamectin, milbemycin oxime, and moxidectin are approved for use. Although MLs have a wide margin of safety, intoxications do occur, especially in dogs (Beal et al., 1999; Hadrick et al., 1995; Hopkins et al., 1990; Hopper et al., 2002; Kenny et al., 2008; Snowden et al., 2006). Collies and Collie-type dogs are at greater risk to develop intoxication due to limited expression of p-glycoprotein activity in the blood–brain barrier (Hopper et al., 2002). There is one case report of intoxication in a kitten (Lewis et al., 1994). A rather unique manifestation of ivermectin intoxication in dogs appears to be blindness as a result of retinal edema

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and electroretinographic changes (Kenny et al., 2008). There are several cases of intoxication in dogs attributed to the ingestion of horse formulations (Beal et al., 1999; Hopkins et al., 1990; Snowden et al., 2006). Chelonians appear to be more sensitive to ivermectin than most mammalian species (Teare and Bush, 1983). Safe doses are up to 10 times lower than those recommended for horses. It is not know whether the increased sensitivity is due to increased permeability of the blood–brain barrier or perhaps due to a higher dependence on peripheral GABA neurons. Intoxication of livestock is relatively uncommon. Adverse reactions to ivermectin in horses appear to be relatively high, however. A survey of 13 Louisiana equine practices reported an approximate 11% incidence of minor to moderate adverse effects (Karns and Luther, 1984). More than 90% of the adverse effects were reported as ventral midline pruritis or edema. A small incidence of injection site swelling, limb edema, eyelid edema, fever, tachypnea, and disorientation was reported. Intoxication of horses following the oral administration of a single dose of 1.87% ivermectin paste has been reported (Swor et al., 2007). Maximum residue levels and withholding times in edible tissues have been established for MLs when used in sheep and cattle (Lanusse et al., 2009). As previously mentioned, MLs are highly lipophilic compounds and are widely distributed in the body. The use of injectable ML formulations in animals that produce milk for human consumption is contraindicated due to residue concerns. However, pour-on and topical formulations of eprinomectin, doramectin, and ivermectin are approved for use in lactating dairy animals. Tissue withdrawal times associated with the use of 1% injectable formulations in cattle for abamectin, ivermectin, doramectin, and moxidectin range from 35 to 50 days (Lanusse et al., 2009). The use of MLs in livestock is of concern from an ecotoxicologic standpoint. The persistent presence of MLs in the feces of treated cattle produces an adverse effect against invertebrates that are important for dung degradation and nutrient recycling to the soil (Floate, 2006). The degree of effect varies among MLs, their formulation, and susceptibility of the insect species (Floate et al., 2005). Long-term environmental consequences of fecal residues of MLs are uncertain.

7.9.5  Neonicotinoids: Imidacloprid   and Nitenpyram Imidicloprid, a chloronicotinyl nitroquanidine compound, was introduced into the United States in 1994 as a veterinary flea control treatment, structural pest and crop insecticide, and seed treatment. For flea control, it is currently marketed as a 9.1% wt/wt solution for use on dogs and cats. Nitenpyram is an orally administered adulticide approved for use on dogs and cats.

Chapter | 7  Pesticide Use and Associated Morbidity and Mortality in Veterinary Medicine

Neonicotinoids for veterinary use are considered to have a low order of toxicity for domestic animals (Hovda and Hooser, 2002). Selectivity is due to differential selectivity for insect nicotinic receptor subtypes compared to mammalian nicotinic receptors (Ensley, 2007). Also, neonicotinoids such as imidacloprid do not readily cross the mammalian blood–brain barrier. There is a single case report of a cat developing dermatosis and associated clinical signs soon after being treated topically with imidacloprid, although no nicotinic signs were noted and another cause for the dermatosis might have been responsible for the clinical signs (Godfrey, 1999). Products containing imidicloprid should not be used on puppies or kittens younger than 4 months of age. There are no reports of nitenpyram intoxication.

7.9.6  Fipronil Fipronil, an N-phenylpyrazole, was introduced into the United States in 1996 for use in animal health, indoor pest control, and commercial turf and crop protection. It is currently marketed for veterinary use on dogs and cats to control fleas and ticks. It is available as a spray (0.29%) and spot-on (9.7 % wt/wt) (Hovda and Hooser, 2002). Fipronil is believed to act as a noncompetitive blocker of GABAgated chloride channels. Veterinary products containing fipronil have a low order of toxicity by dermal, oral, or inhalational exposure for dogs and cats. Fipronil exhibits greater selective toxicity to insects compared to mammals due to GABA receptor affinity differences (Hainzl et al., 1998). However, intoxication can occur due to accidental ingestion or licking of the veterinary product (Gupta, 2007a). Application of the veterinary spot-on can cause skin irritation or hair loss at the site of application (Gupta, 2007a). There is some indication that dogs might be more sensitive to fipronil compared to cats. Off-label use of fipronil in young or small rabbits has been associated with anorexia, lethargy, convulsions, and death (Webster, 1999).

7.9.7  Amitraz Amitraz, an acaricide used to control ectoparasites in animals, has complex pharmacological activity, including 2 adrenergic agonist action. In veterinary medicine, amitraz is used in tick-control collars for dogs at 9%, as a 19.9% topical solution for use on dogs, and as a 0.025% tick dip for use on cattle and sheep (Gupta, 2007b); it is also used on swine. Amitraz is not approved for use in cats, Chihuahuas, pregnant or nursing bitches, or puppies younger than 3 months of age (Baynes, 2009). Accidental consumption of amitraz-impregnated flea collars has caused intoxication of dogs (Grossman et al., 1993). An oral LD50 in dogs is 100 mg/kg; a collar for a large-size dog contains 2.4 g of

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amitraz (Hugnet et al., 1996). Thus, one collar contains an LD50 dose for a 24-kg dog. Amitraz is not recommended for use on horses due to adverse effects. Two cases of colonic obstipation in ponies, probably as a sequel to treatment with an amitraz formulation, have been reported (Mutsaers and van-der-Velden, 1988). Sickness also occurred in three of four horses within 24 h of being sprayed with an 0.025% w/v aqueous suspension of amitraz (Auer et al., 1984). The application consisted of a portion of an amitraz aqueous suspension made up approximately 3 weeks previously to which some freshly prepared spray had been added. It seemed likely that the amitraz in the older solution had broken down to the highly toxic N-3,5-dimethylphenyl N-methyl formamidine derivative and that this was in fact the main cause of the untoward effects observed. The horses displayed typical clinical signs of tranquilization, depression, ataxia, muscular incoordination, and impaction colic lasting up to 6 days. Subcutaneous edema of the face occurred in one horse. Mild dehydration and acidosis accompanied the syndrome. All horses survived after persistent symptomatic treatment, including intravenous fluids, enemas, analgesics every 3 h, multiple doses of paraffin oil per os, and dexamethasone intravenously. Following the eventual relief of the constipation, the horses scoured profusely for 24 h before their condition returned to normal. The reason for the sensitivity of horses to amitraz is unknown, but it is hypothesized to be due to alterations of motility in the large colon resulting in excessive retention of ingestion (Blikslager and Jones, 2004).

7.9.8  Insect Growth Regulators Insect growth regulators (IGRs) first appeared on the market in the 1980s and 1990s and were popular because they were marketed as being harmless to pets, livestock, and humans (Baynes, 2009). Their safety is borne out by LD50s in rodents ranging from 2 to 10 g/kg. IGRs only affect the developing stages of insects and arthropods, not the adults. Therefore, their effectiveness for pest control is not achieved for several weeks after treatment. As a result, many IGRs are combined with adulticidal pesticides such as pyrethrins and pyrethroids or fipronil. IGRs are categorized as either juvenile hormone analogs (approved compounds include methoprene, pyriproxifen, fenoxycarb, and cyromazine) or insect development inhibitors (approved compounds include lufenuron). There are no case reports of intoxication in pets. For lufenuron, doses of 10 and 20 times recommended doses had no adverse effects on cats and dogs, respectively (Hovda and Hooser, 2002).

7.9.9  Synergists and Repellants Synergists such as piperonyl butoxide and N-octyl bicyloheptene dicarboximide (MGK 264) are often included

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in topical insecticide formulations to enhance insecticidal activity. Synergists are less toxic than the active ingredients in topical formulations. When cats are exposed to pyrethrins in combination with piperonyl butoxide at concentrations greater than 1.5%, toxicity of the pyrethrins can be enhanced (Baynes, 2009). Diethyl-m-toluamide (DEET) is approved for use around cats, dog, and horse living and sleeping quarters (Baynes, 2009). DEET concentrations range from 4 to 100%. DEET is not approved for direct application to animals, and there are reports of adverse effects in cats (Dorman et al., 1990). Another repellant, butoxypolypropylene glycol, is formulated with pyrethroids such as resmethrin and permethrin or with piperonyl butoxide in products approved for use in cats, dogs, and horses (Baynes, 2009). No adverse effects have been reported in animals from such products.

There are several reports of cholecalciferol intoxication of dogs and cats (Dougherty et al., 1990; Gunther et al., 1988; Moore et al., 1988; Peterson et al., 1991; Scheftel et al., 1991). Bromethalin intoxication of animals is uncom­ mon but has been reported (Martin and Johnson, 1989). Zinc phosphide intoxication of domestic animals is also uncommon, although there are reports involving dogs, a horse, and chickens (Drolet et al., 1996; Stowe et al., 1978; Tiwary et al., 2005). Cholecalciferol and bromethalin are not restricted-use rodenticides, whereas the availability of zinc phosphide formulations depends on the concentration of active ingredient in a product (products containing 2–5% zinc phosphide are available over-the-counter). Strychnine is a restricted-use pesticide. Strychnine intoxication of domestic animals is not as common as in the past, although cases continue to occur with regularity (Blakley, 1984; Meiser and Hagedorn, 2002; Vig and Dalvi, 1984).

7.9.10  Rodenticides

7.9.11  Metaldehyde

There are a variety of rodenticides used in the environment of domestic animals, including the anticoagulant rodenticides (ARs), cholecalciferol, bromethalin, strychnine, and zinc phosphide. All have been associated with intoxication of nontarget domestic and wild animal species. Currently, the ARs most commonly used in the United States include brodifacoum, bromodiolone, diphacinone, chlorophacinone, and difethialone. They are readily available in a number of formulations, including pellets, paraffin blocks, and tracking powders (Murphy, 2002). Overwhel­ mingly, dogs are the most commonly AR-poisoned species. All of the calls received by the Pet Poison Hotline since 2004 concerning ingestion of long-acting ARs involved dogs (Hovda, 2009). In a survey of practicing veterinarians, AR exposure ranked as the most common toxicant exposure that presented to veterinary clinics (Hall, 2009). Of the currently marketed ARs, brodifacoum is the most commonly implicated in dog poisonings; this most likely reflects the high degree of its use compared to the other ARs. The high palatability of many AR formulations, the indiscriminant eating habits of dogs, and the failure of pet owners to place the products in inaccessible places likely account for this species being most commonly affected. Based on rather limited toxicity information and clinical experiences, cats appear to be less sensitive to ARs. Although relatively uncommon, domestic livestock species can be intoxicated by ARs. There are two reports of chlorophacinone exposure causing death due to acute fatal hemorrhage in cattle and lambs (Braselton et al., 1992; Del Piero and Poppenga, 2006). Anecdotally, horses have also been intoxicated. The rather infrequent poisoning of livestock is most likely due to their relatively large size and the corresponding amount of AR that would have to be ingested to cause intoxication compared to amounts typically used.

Metaldehyde is commonly used in coastal regions as a molluscicide and is available over-the-counter. Formulations include pellets, granules, liquids, or wettable powders (Talcott, 2004). Most products contain less than 5% active ingredient. Metaldehyde is toxic to all domestic animal species, with reports of intoxication of dogs, cats, birds, horses, sheep, swine, goats, and cattle. Dogs are the domestic animal species most often intoxicated (Andreasen, 1993; Talcott, 2004; Yas-Natan et al., 2007).

7.9.12  Paraquat Paraquat is a restricted-use herbicide that is extremely toxic to companion animals and livestock when ingested. Accidental poisoning of animals is uncommon. There is a report of 11 heifers being moderately poisoned by paraquat sprayed on grass along a ditch beside which the heifers walked on their way to and from pasture (Piskac and Jordan, 1970). Dogs have been maliciously poisoned by paraquat. One incident of malicious poisoning involved six dogs in Oklahoma (Bischoff et al., 1998). All the dogs were from the same geographic area, and five had pulmonary and renal lesions consistent with paraquat toxicosis. Another report involved the presumed malicious poisoning of a number of dogs that visited a local park in Portland, Oregon (Cope et al., 2004; Shuler et al., 2004).

7.10  Diagnosis of intoxication Of major difficulty is the early and accurate diagnosis of the specific etiology (i.e., chemical) of an animal mortality event. Circumstances surrounding the toxicosis may direct

Chapter | 7  Pesticide Use and Associated Morbidity and Mortality in Veterinary Medicine

attention to a recent chemical exposure or the introduction of a new animal feed. In some instances, the history of exposure is misleading and diverts attention from subtler and less obvious exposure. Astute evaluation of the clinical signs, prompt postmortem examinations, ascertaining when a herd becomes involved and removal from the inciting source, and then observation of the animals’ responses to therapy can be highly successful in reducing the losses (Oehme, 1987b,c). In many toxicoses, the only clinical sign is death. A thorough postmortem examination is essential in such circumstances. This may help eliminate nontoxicologic etiologies or perhaps narrow the list of possible toxicants. It should be kept in mind that many pesticides might cause nonspecific lesions or no lesions at all. Often, when a postmortem examination is done in the clinic, tissue samples are collected for either histologic or toxicologic examination, but not both. Two sets of tissue samples from animals with suspected toxicoses should be routinely saved. One set should be preserved in 10% buffered formalin for histologic evaluation and another set frozen for possible toxicologic analysis. A common and often unforgiving mistake is failure to submit brain, spinal cord, or peripheral nervous tissue when signs referable to the central or peripheral nervous system are present. A prudent and cost-effective procedure in cases of unexplained deaths is to submit a full set of tissues for histologic examination following gross examination and to keep a second set frozen for later toxicologic analysis pending the histologic findings. It is always easier to dispose of unneeded frozen tissues than to collect tissues from an animal that has already been buried or otherwise discarded. Ideally, a complete postmortem examination should be conducted by a board-certified pathologist at an accredited veterinary diagnostic laboratory with toxicologic testing capabilities. A list of accredited veterinary diagnostic laboratories can be found on the American Association of Veterinary Laboratory Diagnostician’s website at http://www. aavld.org/mc/page.do. Links to individual laboratories are provided. Some laboratories accept only certain animal species for evaluation, and not all laboratories have toxicologic testing capabilities. Therefore, it is useful to call the laboratory for specific information relating to the case at hand. Depending on the toxicant of interest, samples should be either refrigerated or frozen. Refrigeration is generally sufficient if samples are to be shipped and tested soon after collection. If samples are to be retained pending other testing, freezing is more appropriate. When shipping samples, leakproof packaging and appropriate cooling are required. With a few possible exceptions, using ice packs provides sufficient cooling. Samples should be packaged in a way that avoids cross-contamination. Concentrated chemical sources such as baits or products can potentially leak through plastic bags; in such cases, wrapping the sample in aluminum foil or placing it in a separate metal container

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may be warranted. Use overnight delivery services, but avoid shipping samples to arrive on weekends or holidays because the package may sit for a period of time prior to processing. Follow all applicable regulations regarding shipment of biological samples. The sophistication of analytical systems has enabled the diagnostic toxicologist to search for literally thousands of chemical compounds with a high degree of specificity and sensitivity. In cases in which there is no history of exposure to specific chemicals, powerful screening tools such as gas chromatography–mass spectrometry and inductively coupled plasma atomic emission spectroscopy (ICP-AES) are available to detect a broad array of organic and inorganic compounds, respectively. For example, some diagnostic laboratories, using ICP-AES or another metal screening technique called inductively coupled plasma mass spectrometry, can test for several dozen metals with one analysis in tissue, fluid, and environmental samples. Sample size and type can be the limiting factors in the ability of a laboratory to test for a large number of compounds. Small sample sizes limit the number of tests that can be performed and can also decrease the sensitivity of a particular analytical technique. For example, if an analytical procedure typically requires 1 g of tissue to maximize sensitivity, the availability of only 0.5 g can decrease the sensitivity of the test by a factor of two. It is important to note that for certain toxicants, it is not necessary to quantify tissue concentrations; the presence of detectable toxicant in tissues along with compatible clinical signs is sufficient to yield a diagnosis. Alternatively, for agents that are ubiquitous in the environment, quantification of tissue concentrations may be critical for proper differentiation of a toxicosis from a background exposure. An important and potentially vital legal action is securing appropriate tissues and environmental samples for definitive toxicological assay. Although history of exposure, clinical signs, postmortem lesions, and results of therapy are highly significant, the definitive proof of cause is often the identification and quantification of a specific pesticide at toxicological concentrations in the appropriate biological or environmental samples (Oehme, 1999).

7.11  Treatment of intoxication The treatment and management of pesticide intoxications vary with the species of animal involved and the type of offending pesticide. The most important measure is to eliminate or limit further exposure. This involves removal from the offending substance or removal from the offending environment. If full absorption has not occurred, timely decontamination of the gastrointestinal tract or the skin is necessary to limit further absorption. Emetics can be used in small animals, but these are not fully effective and cannot be used in horses and large ruminants due to

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their inability to vomit. Adsorbents such as activated charcoal are clinically useful in preventing the absorption of most organic pesticides from the stomach, but the volume of the rumen makes these adsorbants less efficient in ruminants. Cathartics can be employed to flush out the offending pesticide from the digestive tract, but the large volume and length of the ruminant gastrointestinal system make these treatment procedures less effective than in other animal species. It is generally recommended that gastrointestinal decontamination be performed within 1 or 2 h of exposure. However, the time between exposure and veterinary attention is often greater than this (Cope et al., 2006). Appropriate safety precautions need to be taken by veterinarians and veterinary staff during decontamination procedures to avoid undue exposure. We are aware of a case in which adverse effects occurred in veterinary staff exposed to phosphene gas following a gastric evacuation attempt in a dog that had ingested zinc phosphide. Supportive treatments and emergency intervention are a vital part of pesticide exposure management. In most of the clinically affected cases, vital physiological functions are compromised and life has to be supported until these functions return to normal. In the case of small animals and horses, and to a lesser extent cattle because of cost and availability, specific antidotal procedures can be employed. In veterinary medicine, economics and the willingness of the owner to pay for these potentially beneficial treatment procedures often govern their application.

Conclusion A variety of pesticides are used commonly on animals or in their environment to control a number of economically or aesthetically important external and internal pests. Such use can result in acute or, less commonly, chronic intoxication of animals. The likelihood of intoxication depends on a variety of physiological, behavioral, and environmental factors. Fortunately, the emergence of less toxic pesticides for veterinary use has resulted in less frequent acute animal intoxications. Chronic exposure to pesticides applied to lawns has been hypothesized as a cause of bladder cancer in certain dog breeds, although this remains controversial. In addition to adverse effects on animals, humans can be exposed directly to a pesticide during animal application or indirectly through contact with an animal post application. Another potential route for human exposure to pesticides used in veterinary medicine is via residues in milk, meat, or eggs. Veterinary use of pesticides can have significant environmental impacts through effects on beneficial insects or via secondary intoxication of wildlife. The diagnosis of pesticide intoxication requires careful antemortem and postmortem investigation. Treatment of intoxicated animals involves early decontamination, symptomatic and supportive care, and, in some cases, antidote administration.

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Jaggy, A., and Oliver, J. E. (1990). Chlorpyrifos toxicosis in two cats. J. Vet. Intern. Med. 4, 135–139. Kaluzienski, M. (2000). Partial paralysis and altered behavior in dogs treated with melaleuca oil. J. Toxicol. Clin. Toxicol. 38, 518. Kaneene, J. B., and Miller, R. (1999). Re-analysis of 2,4-D use and the occurrence of canine malignant lymphoma. Vet. Hum. Toxicol. 41, 164–170. Karns, P. A., and Luther, D. G. (1984). A survey of adverse effects associated with ivermectin use in Louisiana horses. J. Am. Vet. Med. Assoc. 185, 782–783. Kendall, R. J., Brewer, L. W., Hitchcock, R. R., and Mayer, J. R. (1992). American pigeon mortality associated with turf application of diazinon AG500. J. Wildlife Dis. 28, 263–267. Kenny, P. J., Vernau, K. M., Pushcner, B., and Maggs, D. J. (2008). Retinopathy associated with ivermectin toxicosis in two dogs. J. Am. Vet. Med. Assoc. 233, 279–284. Kurtz, D. A., and Hutchinson, L. (1982). Fonofos toxicosis in dairy cows: an accidental poisoning (1977). Am. J. Vet. Res. 43, 1672–1674. Lanusse, C. E., Lifschitz, A. L., and Imperiale, F. A. (2009). Macrocyclic lactones: endectocide compounds. In “Veterinary Pharmacology and Therapeutics” (J. E. Riviere and M. G. Papich, eds.), 9th ed., pp. 1119–1144. Wiley-Blackwell, Ames, IA. Lee, L. A., Budgin, J. B., and Mauldin, E. A. (2002). Acute necrotizing dermatitis and septicemia after application of a d-limonene-based insecticidal shampoo in a cat. J. Am. Vet. Med. Assoc. 15, 258–262. Lewis, D. T., Merchant, S. R., and Neer, T. M. (1994). Ivermectin toxicosis in a kitten. J. Am. Vet. Med. Assoc. 205, 584–585. Martin, T., and Johnson, B. (1989). A suspected case of bromethalin toxicity in a domestic cat. Vet. Hum. Toxicol. 31, 239–240. Martínez-Haro, M., Mateo, R., Guitart, R., Soler-Rodríguez, F., Péréz-Lopéz, M., María-Mojica, P., and García-Fernández, A. J. (2008). Relationship of the toxicity of pesticide formulations and their commercial restrictions with the frequency of animal poisonings. Ecotoxicol. Environ. Safe. 69, 396–402. McCurnin, D. M. (1969). Malathion intoxication in military scout dogs. J. Am. Vet. Med. Assoc. 155, 1359–1363. Meiser, H., and Hagedorn, H.-W. (2002). Atypical time course of clinical signs in a dog poisoned by strychnine. Vet. Rec. 151, 21–24. Mensching, D., and Volmer, P. A. (2007). Neurotoxicity. In “Veterinary Toxicology: Basic and Clinical Principles” (R. C. Gupta, ed.), pp. 129–144. Elsevier, Amsterdam. Meyer, E. K. (1999). Toxicosis in cats erroneously treated with 45 to 65% permethrin products. J. Am. Vet. Med. Assoc. 215, 198–203. Mineau, P. (2005). Direct losses of birds to pesticides – Beginnings of a quantification. In “Bird Conservation Implementation and Integration in the Americas: Third International Partners in Flight Conference 2002” (C. J. Ralph and T. D. Rich, eds.), Vol. 2, pp. 1065–1070. U.S. Department of Agriculture, Albany, CA. Moore, F. M., Kudish, M., Richter, K., and Faggella, A. (1988). Hypercalcemia associated with rodenticide poisoning in three cats. J. Am. Vet. Med. Assoc. 193, 1099–1100. Morgan, M. K., Stout, D. M., Jones, P. A., and Barr, D. B. (2008). An observational study of the potential for human exposures to pet-borne diazinon residues following lawn applications. Environ. Res. 107, 336–342. Murphy, M. J. (2002). Rodenticides. Vet. Clin. Small Anim. 32, 469–484. Mutsaers, C. W., and van-der-Velden, M. A. (1988). 2 cases of colonic obstipation in ponies, probably as a sequela of a treatment with Taktic. Tijdschr. Diergeneeskd. 22, 1246–1248 [In Dutch; Abstract only]. National Research Council (1991). “Animals as Sentinels of Environmental Health Hazards.” National Academy Press, Washington, DC.

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Neff, M. W., Robertson, K. R., Wong, A. K., Safra, N., Broman, K. W., Slatkin, M., Mealey, K. L., and Pederson, N. C. (2004). Breed distribution and history of canine mdr1-1Δ, a pharmacogenetic mutation that marks the emergence of breeds from the collie lineage. Proc. Natl. Acad. Sci. USA 32, 11725–11730. Oehme, F. W. (1987a). Anatomical and physiological considerations in species selection – Animal comparisons. In “Human Risk Assessment: The Role of Animal Selection and Extrapolation” (M. V. Roloff, ed.), pp. 47–63. Taylor & Francis, London. Oehme, F. W.; Postgraduate Committee in Veterinary Science (1987b). Agricultural pesticide toxicity in domestic animals. In “Veterinary Clinical Toxicology,” Vol. 103, pp. 429–440. University of Sydney, Sydney, Australia. Oehme, F. W.; Postgraduate Committee in Veterinary Science (1987c). Investigation principles in suspected toxicosis. In “Veterinary Clinical Toxicology,” Vol. 103, pp. 509–519. University of Sydney, Sydney, Australia. Oehme, F. W., (1999). Public health considerations in the use of pesticides and other agricultural chemicals. In “Proceedings of the 26th World Veterinary Congress, Lyon, France” (CD-ROM). Oehme, F. W., and Barrett, D. S. (1986). Veterinary gastrointestinal toxicology. In “Gastrointestinal Toxicology” (K. Rozman and O. Hanninen, eds.), pp. 464–513. Elsevier, Amsterdam. Oehme, F. W., and Rumbeiha, W. K. (1999). Veterinary toxicology. In “General and Applied Toxicology” (B. Ballantyne, T. Marrs, and T. Syversen, eds.), 2nd ed., pp. 1509–1526. Macmillan, London. Osweiler, G. D., Carson, T. L., Buck, W. B., and Van Gelder, G. A. (1985). “Clinical and Diagnostic Veterinary Toxicology,” 3rd ed., pp. 298–320. Kendall Hunt, Dubuque, IA. Perdrizet, J. A., Cummings, J. F., and deLahunta, A. (1985). Presumptive organophosphate-induced delayed neurotoxicity in a paralyzed bull. Cornell Vet. 75, 401–410. Peterson, E. N., Kirby, R., Sommer, M., and Bovee, K. C. (1991). Cholecalciferol rodenticide intoxication in a cat. J. Am. Vet. Med. Assoc. 199, 904–906. Phillips, M. L. (2006). Registering skepticism: does the EPA’s pesticide review protect children? Environ. Health Perspect. 114, A593–A595. Piskac, A., and Jordan, V. (1970). Acute cattle poisoning due to the herbicide Gramaxone. Veterinarstvi 20, 471–473 [In Hungarian]. Post, L. O. (2009). Reporting an adverse drug reaction to the Food and Drug Administration. In “Current Veterinary Therapy, XIV” (J. D. Bonagura and D. C. Twedt, eds.), pp. 99–105. Elsevier, St. Louis. Postgraduate Committee in Veterinary Science (1987). “Veterinary Clinical Toxicology.” University of Sydney, Sydney, Australia. Powers, K. A., Hooser, S. B., Sundberg, J. P., and Beasley, V. R. (1988). An evaluation of the acute toxicity of an insecticidal spray containing linalool, d-limonene, and piperonyl butoxide applied topically to domestic cats. Vet. Hum. Toxicol. 30, 206–210. Raghavan, M., Knapp, D., Dawson, M. H., Bonney, P. L., and Glickman, L. T. (2004). Topical flea and tick pesticides and the risk of transitional cell carcinoma of the urinary bladder in Scottish terriers. J. Am. Vet. Med. Assoc. 225, 389–394. Richardson, J. A. (2000). Permethrin spot-on toxicosis in cats. J. Vet. Emerg. Crit. Care. 10, 103–105. Riviere, J. E., and Sundloff, S. F. (2009). Chemical residues in tissues of food animals. In “Veterinary Pharmacology and Therapeutics,” 9th ed., pp. 1453–1462. Wiley-Blackwell, Ames, IA. Roe, R. T. (1969). The toxicity to cattle of some acaricides in use in plunge dips in New South Wales. Aust. Vet. J. 45, 332–333.

Chapter | 7

Pesticide Use and Associated Morbidity and Mortality in Veterinary Medicine

Royer, H., and Buronfosse, F. (1998). Epidémiologie descriptive des intoxications chez les ruminants (données du CNITV de Lyon de janvier 1990 á août 1998). Le Point Vétérinaire 29, 25–29. Scheftel, J., Setzer, S., Walser, M., Pertile, T., Hegstad, R. L., Felice, L. J., and Murphy, M. J. (1991). Elevated 25-hydroxy and normal 1,25dihyroxy cholecalciferol serum concentrations in a successfully-treated case of vitamin D3 toxicosis in a dog. Vet. Hum. Toxicol. 33, 345–348. Shuler, C. M., DeBess, E. E., Scott, M., and Stone, D. (2004). Retrospective case series of suspected intentional paraquat poisonings: diagnostic findings and risk factors for death. Vet. Hum. Toxicol. 46, 313–314. Senanayake, N. (1998). Organophosphorus insecticide poisoning. Ceylon. Med. J. 43, 22–29. Senanayake, N., and Karalliedde, L. (1987). Neurotoxic effects of organophosphorus insecticides. An intermediate syndrome. N. Engl. J. Med. 26, 761–763. Snowden, N. J., Helyar, C. V., Platt, S. R., and Penderis, J. (2006). Clinical presentation and management of moxidectin toxicity in two dogs. J. Small Anim. Pract. 47, 620–624. Somogyi, A., van Schothorst, M., van Leusdan, F. M., Nouws, J. F. M., Sphon, J. A., Hoffman, B., and Markus, J. R. (1978). Symposium on drug residues in animals tissue. J. Assoc. Off. Anal. Chem. 61, 1182–1298. Stowe, C. M., Nelson, R., Werdin, R., Fangmann, G., Fredrick, P., Weaver, G., and Arendt, T. D. (1978). Zinc phosphide poisoning in dogs. J. Am. Vet. Med. Assoc. 173, 270. Sudekum, M., Poppenga, R. H., Raju, N., and Braselton, W. E. Jr. (1992). Pennyroyal oil toxicosis in a dog. J. Am. Vet. Med. Assoc. 200, 817–818. Sundloff, S. F. (2009). Legal control of veterinary drugs. In “Veterinary Pharmacology and Therapeutics,” 9th ed., pp. 1355–1364. WileyBlackwell, Ames, IA.

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Swor, T. M., Whittenburg, J., Chaffin, M. K., and Barr, C. (2007). Ivermectin toxicosis in horses: 3 cases. J. Vet. Emerg. Crit. Care 17, S16. Talcott, P. A. (2004). Metaldehyde. In “Clinical Veterinary Toxicology” (K. H. Plumlee, ed.), pp. 182–183. Mosby, St. Louis. Teare, J. A., and Bush, M. (1983). Toxicity and efficacy of ivermectin to chelonians. J. Am. Vet. Med. Assoc. 183, 1195–1197. Tiwary, A. K., Puschner, B., Charlton, B. R., and Filigenzi, M. S. (2005). Diagnosis of zinc phosphide poisoning in chickens using a new analytical approach. Avian Dis. 49, 288–291. Vig, M. M., and Dalvi, R. R. (1984). Strychnine poisoning in a dog. Comp. Cont. Edu. Pract. 6, 519–522. Villar, D., Knight, M. J., Hansen, S. R., and Buck, W. B. (1994). Toxicity of melaleuca oil and related essential oils applied topically on dogs and cats. Vet. Hum. Toxicol. 36, 139–142. Volmer, P. A. (2004). Pyrethrins and pyrethroids. In “Clinical Veterinary Toxicology” (K. H. Plumlee, ed.), pp. 188–190. Mosby, St. Louis. Wardhaugh, K. G., Longstaff, B. C., and Lacey, M. J. (1998). Effects of residues of deltamethrin in cattle feces on the development of three species of dung-breeding insect. Aust. Vet. J. 76, 273–280. Webster, M. (1999). Product warning. Frontline. Aust. Vet. J. 77, 202. Wollen, B. H., Marsh, J. R., Laird, W. J. D., and Lesser, J. E. (1992). The metabolism of cypermethrin in man—Differences in urinary metabolite profiles following oral and dermal application. Xenobiotica 22, 983–991. Xavier, F. G., and Kogika, M. M.de S. (2002). Common causes of poisoning in dogs and cats in a Brazilian veterinary teaching hospital from 1998 to 2000. Vet. Hum. Toxicol. 44, 115–116. Yas-Natan, E., Segev, G., and Aroch, I. (2007). Clinical, neurological and clinicopathological signs, treatment and outcome of metaldehyde intoxication in 18 dogs. J. Small Anim. Pract. 48, 438–443.

Chapter 8

Pesticide Use Practices in Integrated Pest Management Frank G. Zalom University of California, Davis, California

Integrated pest management (IPM) has become broadly accepted as an approach to effectively manage insect, disease, nematode, weed, and vertebrate pests in many areas of the world. However, the complexity of IPM is not always appreciated, so the term is often misrepresented. IPM represents the management of pests in a systems framework rather than being simply a tactic or group of tactics for a specific pest or pest group. Many IPM tactics, although they may reduce chemical use, are chemically intensive. The overall goal of IPM is to reduce the environmental and health risks of pesticides within social and economic constraints. IPM has been described as a continuum, with IPM systems ranging from those that are chemically intensive to those that embrace measures that prevent or avoid pest problems and primarily rely on biologically based tactics. A minimum level of IPM requires the use of scouting and decisions based on established action thresholds. Medium-level IPM shifts the management approach largely to preventative measures and relies to a greater degree on the effects of beneficial organisms that are conserved by the avoidance of more broad-spectrum pesticides. High-level IPM systems manage pests through management of ecological and biological processes, often considering the landscape in which the crop or managed site occurs. Moving IPM along the continuum toward the use of more biologically based methods of managing pests remains a challenge that requires extensive interaction between scientists, IPM practitioners, growers, and regulators to ensure relevant development and effective implementation of increasingly more complex IPM systems intended to reduce pesticide risks. IPM practitioners, certified crop consultants, play an especially important role in this process, although pesticides remain one of the primarily curative tools available to them. A wide array of pests, including insects, mites, weeds, nematodes, disease-causing organisms, and vertebrates, Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

lower the quality and the yield of agricultural products, affect the health of humans and other animals, invade structures and landscapes, and adversely affect natural ecosystems. Managing pests has always been a challenge. Before the introduction of synthetic organic pesticides in the 1940s, which allowed reduction of pest abundance and pest damage to levels that were not previously possible, farmers and others responsible for pest control typically employed multiple tactics, such as sanitation, crop rotation, crop diversity, bait trapping, and mechanical pest or host removal, which were applied preventatively based on knowledge of pest biology. Weeds were removed by hand hoeing and tillage; chemical herbicides were seldom applied. They also used inorganic materials, such as copper, lead, antimony, and arsenic, or botanical compounds, such as nicotine and pyrethrum, which were available at the time. These materials were toxic and expensive to produce in quantity; therefore, availability was limited. Equipment for their application was relatively unsophisticated or lacking. Overall, pesticide use was low relative to contemporary levels. The chemical control paradigm was developed effectively by industry, government, and university researchers, and it became widely implemented. Along with modern plant breeding, fertilization, and irrigation methods, the introduction of synthetic pesticides reduced on-farm labor requirements, facilitating the transition of agricultural production in developed countries to a highly mechanized system with relatively more concentrated production that is characterized by increased yields and reduced variability in production. Arguably, this transition has been beneficial in that fewer people must work on farms to produce the food and fiber products required to sustain an ever-growing population. The cost of food and fiber remains low as a proportion of income, and food supplies are relatively stable in developed countries. Unfortunately, despite an extensive regulatory system for registration, the increased use of

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pesticides has been accompanied by unintended social and environmental consequences, including documented cases of pest resistance and pesticide-induced pest outbreaks, environmental contamination, worker exposure, and public concern for residues on food. The only way to eliminate the risk of using pesticides is to prohibit their use, but at what cost? Pesticides are legally classified as economic poisons and are defined as substances used to control, prevent, destroy, or mitigate any pest. Pesticides include inorganic products such as sulfur and naturally occurring botanical products such as pyrethrum, both of which are acceptable for use by organic growers. Pesticides include vegetable and petroleum oils, fertilizers, and certain fatty acid soaps when they are used for pest control. Naturally occurring microbes, such as Bacillus thuringiensis and Trichoderma harzinium, are considered pesticides when they are produced commercially and marketed as pest control agents. Many pesticides are very specific in their actions, acting as growth regulators, repellents, pheromones, desiccants, and defoliants. However, the general public seems most concerned with the use of certain synthetic pesticides, particularly those with broader activity and those with which they are unfamiliar. Agricultural uses in particular are not well understood by the public, so questions abound concerning the safety of these products and the need for their use.

8.1  Integrated pest management The idea of “integrated control” is not new. Hoskins et al. (1939), as cited in Smith (1974, p. 427), are believed to have been first to use the term when they commented that biological and chemical control are considered as supplementary to one another or as the two edges of the same sword. … Nature’s own balance provides the major part of the protection that is required for the successful pursuit of agriculture. … Insecticides should be used so as to interfere with natural control of pests as little as possible.

This was before the pesticide era. However, the idea was more broadly accepted in the 1950s, when identification of pest resistance, pesticide-induced pest outbreaks, and the resurgence of pests that had been under control led some researchers (e.g., Michelbacher and Bacon, 1952; Smith and Allen, 1954; van den Bosch and Stern, 1962) to call for integrated control to reflect the combination of compatible biological and chemical control tactics. The concept was expanded to include economic thresholds by Stern et al. (1959), who called their approach integrated pest management or IPM. Economic thresholds are the pest densities at which the value of resulting damage exceeds the cost of applying a control. Their description of IPM added the requirements of pest monitoring and risk assessment before justifying the application of therapeutic measures such as pesticides.

Environmental contamination by organochlorine insecticides was recognized in the 1960s, following the publication of the book Silent Spring by Carson (1962). Pesticide use became a political issue, and IPM was promoted as an acceptable approach for managing agricultural pests among some scientists and growers who were interested in applying “supervised control” rather than using strictly preventative pesticide treatments, which had become prevalent by that time. However, concerns about the slow rate of IPM adoption by farmers were raised by IPM researchers (e.g., van den Bosch, 1964). Funding for IPM research increased greatly during the 1970s and early 1980s, with increasing efforts to implement IPM practices through extension services, governmental agencies, and community-based programs. As the philosophy of IPM matured, there grew an ever greater appreciation for integrating the management of weeds, pathogens, and nematodes as well as insects in a cropping systems context, recognizing that fundamental differences exist in the biology of these pests and, therefore, in the preventative and therapeutic measures that can be applied for their control. IPM strategies and tactics have gradually been adopted as alternatives to the conventional chemical control paradigm, and the breadth of institutions and organizations promoting IPM as the most effective way to reduce the risks of using pesticides has dramatically increased. The United Nations Food and Agriculture Organization’s Panel of Experts on Integrated Pest Control (1967, p. 3) defined IPM as “a pest management system that, in the context of the associated environment and the population dynamics of the pest species, utilizes all suitable techniques and methods in as compatible a manner as possible and maintains the pest populations at levels below those causing economic injury.” In the United States, several administrations have endorsed IPM. The U.S. Department of Agriculture (USDA) Council on Environmental Quality (1972), in its publication Integrated Pest Management, wrote that IPM is an approach that employs a combination of techniques to control the wide variety of potential pests that may threaten crops. It involves maximum reliance on natural pest population controls, along with a combination of techniques that may contribute to suppression-cultural methods, pestspecific diseases, resistant crop varieties, sterile insects, attractants, augmentation of parasites or predators, or chemical pesticides as needed.

In urging IPM adoption in an environmental message, President Carter (1979) said that “IPM uses a systems approach to reduce pest damage to tolerable levels through a variety of techniques, including natural predators and parasites, genetically resistant hosts, environmental modifications, and, when necessary and appropriate, chemical pesticides. IPM strategies generally rely first upon biological defenses against pests before chemically altering

Chapter | 8  Pesticide Use Practices in Integrated Pest Management

the environment.” Attention to IPM in the United States increased again following the Clinton Administration’s 1993 pledge to have 75% of cropland acreage under IPM by the year 2000 and to reduce the use of pesticides. Meeting the year 2000 goal was discussed in a report prepared by the U.S. General Accounting Office (2001). The report cited the National Agricultural Statistics Service (USDA NASS, 1998) as estimating that some level of IPM had been implemented on approximately 70% of the nation’s crop acreage by the end of crop year 2000, but it examined these same data to conclude that the level of biologically based IPM being practiced was relatively low. It also suggested that the increase in IPM use reported by NASS was not reflected in a concomitant decrease in the amount of pesticide used. The USDA responded to the report by pledging to make management of the program a high priority, and the U.S Environmental Protection Agency (U.S. EPA) emphasized that promoting IPM is an important component of the agency’s approach toward reducing risks posed by pesticides. A significant outcome was the development of a national IPM road map (Coble and Ortman, 2004) that identifies strategic directions for IPM research, implementation, and measurement for all pests, in all settings, throughout the nation, including not only agricultural but also structural, ornamental, turf, and human and wildlife health pests. The IPM road map reflected the significant expansion in the scope of IPM that had occurred during the previous decade into all areas in which pests and pesticides are important. Indeed, IPM has become the preferred pest management approach for many municipalities, schools, and parks. Federal departments such as the Department of Defense and Department of Housing and Urban Development have implemented active IPM programs for housing and lands under their jurisdictions. The road map also placed the focus of IPM evaluation not on reducing pesticide use but, rather, on reducing the risks associated with pesticide use, which is an important distinction.

8.2  What is integrated pest manangement? Forty years after the term “integrated pest management” first appeared in the literature, a single definition has yet to be universally adopted. This is not unexpected because IPM can be as much a philosophy as a science. Because IPM has diverse proponents, the term has been adapted to support a variety of objectives and agendas. This adaptation has tended to permit narrow definitions of IPM to be proposed, in which it is mentioned primarily in terms of tactics such as chemical controls or biological controls, which have particularly strong advocates. What was largely promoted as an ecologically based view of pest management by a relatively small group of academics and certain agricultural interests in the 1950s and 1960s has become a term for reaching a useful compromise among environmental groups and those interests that must

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use pesticides to manage damaging pest populations. The recent focus of IPM on reducing pesticide risks as opposed to strictly reducing use has been useful in this regard. Allowing diverse groups to reach common ground is indeed a strength of IPM. However, in many ways it has also become one of its major shortcomings. Depending on its interpretation, IPM can be used to justify conventional pest control practices, even those that are chemical intensive, without emphasizing reduced-risk alternatives or, more important, management of the pest species within an ecosystem framework. Cate and Hinkle (1993) stressed the ecological basis of IPM, rather than the tactical emphasis of many IPM definitions, in their report “Integrated Pest Management: The Path of a Paradigm” by accurately stating that IPM is about the manner by which communities are managed. Perhaps the phrase itself has resulted in misinterpretation. Kogan (1988) identified “integrated” as the most ambiguous component of the term integrated pest management. To many people, “integrated” refers to the use of multiple control tactics integrated into a single pest control strategy (Metcalf and Luckmann, 1982). This strategy most typically targets only one species of pest or a single class of pest and, in this sense, focuses on control measures for the target species, prevention of natural enemy disruption and secondary pest outbreaks, and delaying development of pesticide resistance. A broader interpretation refers to management of the complex of pests that attack a crop, considering the combined effects of weeds, plant diseases, insects, and nematodes (Newsom, 1980). At its highest level, IPM incorporates interactions among pests, the crop, and the environment within the context of a social, political, and economic matrix. Prokopy (1994) likened the increasing levels of IPM complexity – from integration of control methods for a specific pest to its incorporation into a socioeconomic matrix – to the steps of a ladder, where progressing up the steps represents increased levels of integration in a systems context. The word “management” as opposed to “control” also presents an important IPM concept. Flint and van den Bosch (1981) stated that the word “management” implies acceptance of pests as inherent components of an agricultural system. Indeed, some would say that acceptance of pests in an agricultural system is essential to permit their natural enemies to survive in an ecosystem. The IPM approach is to apply controls to suppress pest populations when necessary to reduce damage to an acceptable level rather than to eradicate the pest.

8.3  The IPM continuum IPM systems have been characterized as progressing along a continuum (Sorenson, 1993) ranging from those that are more chemically intensive, where pesticides are applied based on scouting and the use of thresholds, to those that are biologically intensive, where reduced-risk pesticides

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may be applied but biological control, conservation, and biologically based preventative approaches predominate. The USDA formalized the continuum concept in quantifying IPM adoption by creating categories for no IPM use and three additional levels, which represent progressively greater use of biological or cultural practices instead of conventional pesticides (USDA NASS, 1998; Vandeman et al., 1994). When presented as a continuum, the minimum criteria that constitute the use of IPM are field scouting for both pests and natural enemies and using action thresholds where they exist to make pesticide use decisions. When an action is warranted, those people who employ a minimum level of IPM would apply selective or the “reduced-risk” pesticides. A medium level of IPM would include scouting and thresholds and also one or two preventative practices, whereas the USDA IPM continuum required at least three preventative practices. The USDA’s continuum concept was useful at the time because it provided a mechanism to evaluate adoption of specific tactics across many crops, but others have subsequently modified it to specify the use of at least some preventative practices at all levels and to emphasize systems and biological controls (Benbrook et al., 1996; Hoppin et al., 1996; Jacobsen, 1997; Kogan, 1998). At its highest level, a bio-intensive IPM system would include release or conservation of beneficials, trap crops, use of interactive pest and weather/crop models, and other primarily nonchemical preventative approaches.

8.4  Pesticides Although IPM emphasizes a systems approach to management, it is impossible to discuss the practice of managing pests without mentioning tactical intervention. Pesticides often represent the first line of defense in situations of pest outbreak or when a specific pest must be eradicated for quarantine or public health purposes. As mentioned previously, pesticides may be used in an IPM system when applied based on scouting and strict consideration of available action thresholds. However, where choices of pesticides exist, those that are least toxic and present the lowest potential for disruption, so-called reduced-risk pesticides, should be selected for use. Wigglesworth (1950) pointed out that it is sometimes “through the activities of the entomologists themselves that entomological problems arise.” He also stated that the public loves the hospital, the doctor, and the bottle of physic; while the advances in preventative medicine which have transformed our lives are scarcely noticed. So too it creates a greater impression on the mind to destroy an infestation of insects that can be seen, than by some simple change in practice prevent any infestation from developing.

Pest resistance to specific pesticides and pest outbreaks that result from applications of broad-spectrum pesticides are problems frequently associated with overuse or misuse of pesticides, and their occurrences are well documented.

Pesticide resistance is the adaptation of a pest species to a pesticide, resulting in its decreased susceptibility to that chemical. Even when a high percentage of the population is killed by a pesticide application, those few individuals that possess the resistant traits will survive and reproduce, passing their genes to the succeeding generation. Thus, a pest population develops that can be controlled only by higher chemical dosages. After a population is exposed to a pesticide for a prolonged period, it no longer kills the population effectively. Thus, most pesticides have a finite effective life. Pest resistance to a chemical can develop rapidly, particularly when the life cycle of a pest species is relatively short, fitness is high, the treated population is relatively isolated from untreated populations, and the chemical is repeatedly applied. Pesticide resistance has been documented in hundreds of species of insects and mites, plant pathogens, weeds, rodents, and nematodes (Georghiou, 1986). The best way to manage pest resistance is to apply pesticides less frequently as part of a more comprehensive IPM approach. IPM tactics such as scouting and applying pesticides only when needed are basic to any IPM program. However, emphasizing nonchemical tactics, such as beneficial insects (predator/parasites), cultural practices, transgenic plants, crop rotation, pest-resistant crop varieties, and pheromone mating disruption, that may reduce the need to apply pesticides advances the IPM concept. When pesticides must be used, alternating classes of pesticides applied to reduce selection pressure on the pest population can delay the development of resistance. Applying biological insecticides, when available and effective, in rotation with other products is also useful. Approaches for monitoring susceptibility of pest populations to specific pesticides have been developed for several key arthropod and disease species. Such technology is relatively common for research applications and when applied by the pesticide manufacturer, but commercial implementation of resistance monitoring by IPM practitioners is rare. Many plant-feeding insects do not significantly damage agricultural crops because they are kept under natural control by predators and parasites. However, these natural control agents can be inadvertently disrupted by chemical applications that target the bona fide pest species. This situation can result in the emergence of secondary pests, which have been released from natural control. For example, it is widely believed that spider mites have emerged as serious agricultural and forest pests primarily because their predators have been reduced in abundance by chemical sprays for primary pests.

8.5  Field scouting 8.5.1  Monitoring Field scouting or monitoring includes proper identification of pests through surveys or scouting programs, and it may

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incorporate trapping, weather monitoring, and soil testing where appropriate. In many instances, pesticide use for controlling a given pest has been reduced 40% without affecting quality or yield simply by using quantitative monitoring procedures in combination with realistic control action thresholds (National Research Council, 1989). Monitoring may be supported through the use of phenology or risk assessment models or through other types of decision support. In practice, monitoring can be done by either the grower or practitioners who check the fields for growers, but there is a labor cost associated with monitoring that is not associated with the preventative use of pesticides. The challenge for researchers is to develop commercial monitoring plans that are economically implementable as opposed to sampling regimes developed for research purposes. The lack of practical monitoring procedures and use of those procedures results in poor timing of applications and an excessive use of pesticides. University pest management guidelines that are now common on the World Wide Web often offer useful monitoring programs for many pests.

8.5.2  Decision Support One focus of IPM research for many years has been the development of models that present a framework for integrating information from the various biological disciplines, meteorology, and the field monitoring of pest populations. These models have served to bring disciplines together in analyses of production systems and have yielded tools that can be implemented to support the monitoring or scouting process. IPM research has pioneered many applications for computer technology in agriculture and helped to bring about the early use of electronic instruments for field data gathering (Zalom and Strand, 1990) and integrating these data into predictive tools. Insect monitoring, which incorporates the use of bait or pheromone traps, is an approach that has become commonly used for monitoring various pest species, providing information on the mobile adult stage. When used in conjunction with phenological models, monitoring can be used to predict pest development and, ultimately, to accurately time pesticide applications. Dozens of insect phenology models based on degree-day accumulations are now used by IPM practitioners to improve timing of pesticide applications. Advances in technologies for monitoring temperature and leaf wetness led to commercial implementation of risk assessment models for several key diseases, including late blight of potato (Phytophthora infestans) (Krause and Massie, 1975; Stevenson, 1983), grape powdery mildew (Uncinula necator) (Gubler, 1991; Sall, 1980), and Alternaria solani on tomato (Madden et al., 1978). TOMCAST, a tomato disease forecasting program designed to predict early blight, septoria leaf spot, and anthracnose

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(Pitblado, 1992), has become widely available to growers through commercial services, university extension services, and the Campbell Soup Company (Bolkan and Reinert, 1994). Model predictions are usually made by first predicting when conditions are met that are favorable to disease development and then assessing the severity using a disease risk index. Commercial validation of risk assessment models has shown potential for reducing the number of pesticide applications, depending on year, geography, and disease pressure (Weber et al., 1996). In 2005, the Integrated Pest Management Pest Information Platform for Extension and Education (IpmPIPE) was established as a dynamic, integrated national system facilitated by information technology that provides centralized, useful tools with reliable information for IPM practitioners. The first application of the system was to serve as a national warning system designed to help soybean farmers protect their crop from the devastating introduced disease Asian soybean rust; it is accessible via the Internet (http://sbr.ipmpipe.org/ cgi-bin/sbr/public.cgi).

8.6  Reduced-risk pesticides When chemical tactics are deemed necessary in an IPM system, the choice of a product that is selective and least disruptive to the ecosystem and human health is desirable. Some pesticides are by their nature less risky. For example, many biological pesticides that are derived from microbes, plants, and certain minerals pose a lower risk. There has been a great expansion in the availability and use of reduced-risk products, especially since the passage and implementation of the Food Quality Protection Act of 1996. The U.S. EPA has also given priority in its registration program for conventional chemical pesticides to pesticides that meet reduced-risk criteria: low impact on human health, low toxicity to nontarget organisms (birds, fish, and plants), low potential for groundwater contamination, lower use rates, low pest resistance potential, and compatibility with integrated pest management (U.S. EPA, 1997).

8.6.1  Behavioral Chemicals Pheromones are highly specific chemicals released by insects to affect the behavior of members of their own species, usually as attractants for mating but also as signals for aggregation, alarm, or feeding. Synthetically produced pheromones are frequently used in IPM programs as described previously to monitor adult insect flights. The direct use of pheromones as control agents has also met with some success, usually when the chemical is released over the field from dispensers with the intent of confusing males and preventing mating by inhibiting their ability to locate females. This technique has been applied for control of such key pests as the oriental fruit moth [Grapholitha

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molesta (Busck)] in Australia and California (Rice and Kirsch, 1990), the tomato pinworm (Keiferia lycopersicella) in Mexico and the United States (Jimenez et al., 1988), the codling moth (Cydia pomonella) (Brunner, 1994), the pink bollworm (Pectinophora gossypiella) (Flint et al., 1993), and the peach twig borer (Anarsia lineatella) in California (Pickel et al., 2002). Because mating disruption is more effective when deployed over larger areas, areawide programs have been established by the USDA to implement mating disruption for key pests regionally on extensive crop acreage in the United States. The areawide codling moth management program conducted on pears and apples in the Pacific Northwest and California (Calkins and Faust, 2003) was one of the most outstanding examples of such a program, having led to widespread use of mating disruption and other reduced-risk approaches.

8.6.2  Conventional Products and Risk Reduced-risk conventional pesticides are usually considered to be safer than traditional pesticides in terms of toxicity to humans and the environment. The Food Quality Protection Act of 1996 was a major stimulus for the registration of reduced-risk pesticides by agrichemical manufacturers. As regulatory pressures result in decreased use of older classes of pesticides such as organophosphates and carbamates due to restrictions on their use or outright cancellation of uses, those pesticides that appear to have a reduced-risk profile are becoming increasingly available. They are favored in the registration process over products that are anticipated to present greater risks, so more products representing a multitude of novel modes of action are being brought to market by manufacturers. Historically, one type of risk often cited when promoting risk reduction in environmental and health terms was the financial risk associated with using less effective controls. Risk is probably the most important financial obstacle to IPM adoption. Growers value pesticides for reducing production risk as well as contributing to profit. For more biologically intensive IPM systems to be voluntarily adopted, it has often been stated that IPM must be shown to decrease financial risk (Antle and Park, 1986; Gruys, 1982; Way, 1977). In reality, IPM strategies such as monitoring are effective tools for managing risk. The more growers learn about pests and their damage potential under an IPM scenario, the less is the uncertainty in their minds about the state of their crop and the more likely it becomes that they will not choose to make a preventative pesticide application. A number of reduced-risk insecticide products such as growth regulators have been shown to be highly effective substitutes for conventional products, so their adoption is increasing. However, the increased cost of these products relative to conventional products remains a factor restraining use.

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8.7  Cultural and physical suppression Cultural controls have been used historically to manage many pests, but these were often abandoned in favor of pesticides that were less labor-intensive. Such controls include a broad range of production practices that render the crop environment less favorable for the pest. Tillage and water management are effective cultural controls in the management of weeds. Furthermore, increased mortality in many insects that overwinter in the soil may result from particular tillage practices. Narrow row plant spacing or optimal in-row spacing can also suppress weeds under certain cropping systems. The destruction of crop residues is important in the management of many pests, such as navel orangeworm in almond, late blight of potato, stem rot of rice, and pink bollworm and boll weevil in cotton, for which there are compulsory plow-down dates in several regions. Physical suppression tactics may include cultivation or mowing for weed control and temperature management or controlled atmospheres for postharvest pests.

8.8  Prevention Pests are managed in an IPM system in part by preventing their occurrence. Prevention includes those practices that keep pests from invading a crop or field and then becoming established. It includes such tactics as using pest-free seeds or transplants; excluding pests by screens or barriers; preventing weeds from reproducing by disking or mowing; choosing plant cultivars with genetic resistance to insects, nematodes, or diseases, as well as benefits that result from genetically modifying organisms; irrigation scheduling to avoid situations conducive to disease development; cleaning tillage and harvesting equipment when moving between fields; using sanitation procedures to remove an incipient infestation; and eliminating alternate hosts or sites of pest organisms. Even applying fertilizer with the seed of annual crops or through drip irrigation systems can provide a measure of weed control, especially in contrast to broadcast application of fertilizers, which stimulates weed growth.

8.9  Avoidance Avoidance is practiced when pest populations exist in a field or site, but the impact of the pest on the crop can be avoided through some cultural method. Examples of avoidance tactics include crop rotation to break the life cycles of pest species, using trap crops, choosing plant cultivars with maturity dates that may allow harvest before pest populations develop or that have a sufficiently short season to permit planting after conditions are conducive to infestation, fertilization programs to promote rapid crop development, and simply

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not planting certain fields or areas within fields where damaging pest populations are most likely to develop.

8.10  Pesticides and biological controls Biological control is the augmentation, conservation, and importation of natural enemies, including predators, parasites, and pathogens, to reduce a pest population. This may involve either introduction of a natural enemy or augmentation of one that already exists in the crop ecosystem. Biological control is generally considered to be the cornerstone of any IPM program. Mass culture and release of predatory lacewings, various species of parasitic wasps, and insect pathogens such as Bacillus thuringiensis have been effective in certain insect pest management programs. They are especially important in organic production. The release of certain species of plant-feeding insects and pathogenic fungi has been successful in controlling a number of important rangeland weed pests as well. Biological control has the advantage of generally being safe to nontarget organisms, although there is concern that introduced biological control agents be specific so as not to disrupt native systems. Classical biological control, the release of an imported natural enemy to control a pest species, when successfully established remains more stable in the environment than other pest control tactics. Natural enemies used in augmentative releases often do not persist and must be re-released periodically. Conserving natural enemies by avoiding disruptive sprays has become an essential practice in IPM cropping systems. Prior to the 1940s, spider mites were considered to be sporadic pests in most perennial crops. Following the introduction and widespread use of broad-spectrum pesticides, spider mites became annual pests in many crops. One of the best examples of conserving natural enemies through the careful use of pesticides involves apple production in the Pacific Northwest, where the spider mite Tetranychus mcdanieli (McGregor) can be a primary arthropod pest. Beginning in the mid-1960s, an effective mite management approach based on the conservation of the western orchard predator mite, Galendromus occidentalis (Nesbitt), using selective insecticides for control of orchard pests was developed and implemented by Hoyt (1969). In this system, organophosphate cover sprays could be used for the codling moth (C. pomonella), the key pest of apples if an alternate prey, the apple rust mite [Aculus schlechtendali (Nalepa)], was encouraged to support populations of G. occidentalis. Implementation of this program reduced the average mite control cost for Washington state growers from $24 per hectare in 1967 to $8–$12 per hectare in 1985 (Croft, 1990). Ironically, as organophosphates and carbamates of insecticides are being replaced by other products, there

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exists a danger of disruption of IPM systems that have successfully integrated the use of conventional chemical and biological control tactics. The increased use of pyrethroids in many cropping systems presents such a danger because they have been shown to be highly disruptive in orchards by killing predator mites (AliNiazee, 1984; Croft and Hoyt, 1978). Residues of pyrethroids have been shown to be persistent and remain biologically active against predatory mites long past their initial application (Zalom et al., 1998), and there is some indication that the residual effects of pyrethroids persist on orchard trees into the subsequent growing season (Bentley et al., 1987). Often, the effects of pesticides on biocontrol agents are not obvious. Even traditional bioassays used to determine direct mortality do not identify potential problems that might arise from use of a product. Pesticide toxicity has traditionally been evaluated by considering adult female mortality as the endpoint – that is, estimating values that measure median lethal concentration (LC50) or median lethal dose (LD50) (Robertson and Worner, 1990). Because these evaluations focus on a single life stage and generally for a short duration of time (often 1–4 days), the results of these bioassays do not accurately assess the total effects of a pesticide on an exposed population (Stark and Banken, 1999). Evaluation of sublethal effects together with assessment of acute effects to estimate the total effect of a pesticide gives a more complete picture of the risks (Stark et al., 1995). By examining the total effects of several new acaricides on two species of predatory mites, Saenz de Cabazon Irigaray et al. (2007) showed that a presumably reducedrisk mite growth regulator in fact had a catastrophic effect on predator reproduction. More such studies are needed to fully appreciate the risk posed by reduced-risk products on beneficials. The release of natural enemy strains selected for resistance to disruptive insecticides allows the selected natural enemies to persist even when the disruptive materials are applied for control of key pests. A laboratory-selected strain of the predator mite G. occidentalis, resistant to carbaryl and organophosphates, was successfully used to manage spider mites in California almond orchards (Hoy et al., 1984). Extensive research was conducted on their economical mass-rearing (Hoy et al., 1982), sampling (Wilson et al., 1984; Zalom et al., 1984), and applications of selective acaricides at lower than label rates to help adjust the ratio of spider mites to predator mites in favor of the predator mites, thus enabling the integration of this approach with other almond orchard practices.

8.11  Advisory services The practice of IPM required information about the status of pests and their natural enemies in the context of the system being managed. The development of a pest management

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consulting industry can be viewed as one of the most positive results of IPM implementation efforts, and certified IPM practitioners and crop consultants have become a major force in the delivery of pest management information and decision support in agricultural, landscape, and urban settings (Blair, 1986; Frisbie and McWhorter, 1986; Green et al., 2007). Pest management is becoming ever more complex due to changes in pesticide availability and use restrictions in commercial settings such as agriculture and structures and due to public demand for government agencies such as parks departments, public housing authorities, and schools to reduce overall pesticide use in landscape and urban settings. Therefore, growers and agencies increasingly rely on the advice of third parties for their pest management decisions. IPM practitioners and certified consultants have come to play a significant role in implementing IPM in managed systems. This is particularly true when implementation of more biologically intensive or higher level IPM is the goal. References in the scientific literature to crop consultants providing pest management information to their grower clientele date back more than 60 years, and this form of pest management consulting was referred to as “supervised control” (Michelbacher, 1945). The number of individuals practicing supervised control remained relatively small until the early 1970s, when the USDA initiated IPM pilot projects in a number of states to demonstrate the benefits of scouting programs and to promote their use. An analysis of selected extension IPM programs including many of the pilot projects was conducted by Rajotte et al. (1987). To distinguish individuals who practiced higher level pest and crop management, professional certification programs have been initiated. The principal professional certification programs in agricultural crops are the Certified Crop Adviser and the Certified Professional Crop Consultant, with certification from the American Society of Agronomy and the National Alliance of Independent Crop Consultants, respectively. There are more than 13,000 Certified Professional Crop Consultants and 500 Certified Professional Crop Consultants nationally. These programs share similar requirements in that applicants must pass one or more comprehensive examinations, possess a bachelor’s degree and/or documented experience in advising, and agree to sign a code of ethics developed by each organization. Both programs require approved continuing education coursework to maintain certification. IPM certification programs similar to those for crop consultants have been developed for structural pest control operators who wish to be identified as IPM practitioners. These programs, Green Shield Certified and EcoWise Certified, place an emphasis on higher level IPM use rather than routine pesticide applications. A program for licensing pest control advisers was initiated in 1976 by the state of California, and approximately

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3800 individuals are now so licensed. The law requires anyone who recommends pesticides or any other pest control method or device for agricultural use to be licensed. This law has had a far-reaching effect on increasing the number of growers who use a minimum level of IPM; for most crops, a higher proportion of acres are scouted in California than elsewhere in the United States. Like the national certification programs, the licensing process requires passing a comprehensive examination in pesticide laws and regulations and basic IPM principles. It also requires a minimum of a bachelor’s degree in agricultural sciences, biological sciences, or pest management, and a specified number of course units from a list of approved classes. There is a continuing education requirement of 40 h every 2 years. The potential effect of California’s licensing program may not have been realized because the licensing program does not distinguish private consultants from the majority of consultant who work for farm supply dealers or other pesticide retailers (Wearing, 1988). Although California pest control advisers no longer receive commissions or bonuses based on their sales of farm chemicals, an incentive to consider alternative practices, including taking no action, may be lacking. A survey of current employment of all licensees by the California Association of Pest Control Advisors (CAPCA, 2006) indicated that 16% worked for retailers and 15% worked for basic manufacturers, whereas 22% (vs. 14% in 1999) were independent consultants and 17% in-house consultants who worked directly for farming operations. The remainder were largely government or municipal employees. Interestingly, a survey of California almond growers indicated that although growers who used independent pest control advisers tended to feel more knowledgeable about IPM and reported using more complex pest monitoring techniques and control practices, their use of insecticide sprays was independent of the type of employment of their pest control adviser (Brodt et al., 2005). Despite certification and licensing programs for IPM practitioners and consultants, a majority of growers in many areas of the United States practice a form of self-treatment with pesticides, controlling the choice of chemicals and treatment schedules. As long as pesticides are used according to label restrictions, there are few additional restrictions on their availability or use. With few exceptions, there is no requirement that treatments be based on an accurate diagnosis of a problem or whether, in fact, a problem exists. There is no requirement that alternative treatments be considered or that knowledge of alternative treatments exist. This undoubtedly contributes to the public’s negative attitude toward agricultural chemicals. Increasingly more stringent regulations will likely lead to the loss of higher risk pesticides or use of pesticides for which alternatives are less effective or more costly to growers.

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Prescriptive use of human drugs by licensed physicians has not eliminated cases of injury due to their application, the development of resistance to drugs, or other ancillary problems. However, there is public confidence in the regulatory system for medicine and drugs that does not exist in the regulatory system for pesticides. Would the prescriptive use of pesticides by licensed practitioners help to improve public confidence in the use of pesticides? Coble et al. (1998) addressed this issue by proposing a model similar to that used in the medical profession whereby relatively low-risk chemicals may be self-prescribed, but high-risk chemicals may be prescribed only by a trained and licensed professional. This proposal is one mechanism by which certain valuable pesticide uses could be maintained while addressing the public’s concern for safe use of those products. Already, pesticides are not treated equally in the registration process. For example, pesticides that present the greatest risk to human health or the environment have various restrictions placed on their use. Pesticides that are believed to be “safe” may be put on a fast track for registration.

Conclusion There are many challenges to the development and implementation of IPM systems, but an excellent framework exists in the scientific literature and in experiences with successful field implementation. The concept of integration began with the realization that the use of synthetic pesticides, which helped to make pest control more predictable and less labor-intensive, brought about certain unintended consequences, such as pest resistance, secondary pest outbreaks, and the resurgence of pests that previously had been under good control. Integrated control suggested that by utilizing pesticides in such a manner as to preserve naturally occurring biological control, more effective and, in the long term, more economical pest control could be achieved. Integrated pest management incorporated the concept that pesticides should be used only when needed based on careful assessment of the risk posed by specific pest densities and the potential for control of those pests by naturally occurring beneficial organisms or other factors in the environment. IPM became more interdisciplinary, incorporating an ecosystem approach. As concern about the impact of pesticides on the environment and on human health became elevated in society, IPM gained favor in academic and governmental circles as an acceptable strategy for managing pests. With wide acceptance of the paradigm came a particular emphasis on IPM tactics within the range of practices that can be utilized in an IPM system. A number of authors refer to the range of IPM practices as falling along a continuum from those that are more chemically intensive to those that are more biologically intensive, and they have attempted to categorized these into different “levels” of IPM.

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It is clear that many or even a majority of growers practice a minimum level of IPM. However, why has there not been more progress toward implementation of less pesticide-intensive IPM systems? A significant body of literature (Sorenson, 1994; Wearing, 1988; Zalom, 1993) discusses the technical, financial, educational, institutional, and social constraints to IPM use. The National IPM Forum held in Washington, DC, in 1992 identified the lack of a national commitment to IPM, lack of funding for IPM research and extension activities, perceived problems with the regulatory process that affects registration of new technologies, and the shortage of well-trained, independent IPM consultants as major issues in the advancement of IPM. Since the forum, the national commitment to IPM increased with the Clinton Administration’s 1993 pledge to have 75% of cropland acreage under IPM by the year 2000. Also, the U.S. EPA has made significant changes in the way it approaches registrations of new products by establishing a fast track for certain “reduced-risk” compounds, and the agricultural chemical industry has responded with a number of new products. However, funding for IPM research and extension activities, particularly those that emphasize integrative and ecological approaches, has not occurred, and relatively minimal incentives have become available for the adoption of IPM practices. Training of a new generation of IPM practitioners will perhaps be the major challenge to advancement of IPM in the next decade as agricultural colleges decrease emphasis on IPM training and the demographics of IPM practitioners show a steady increase in age. For example, 40% of licensed pest control advisers in California are older than age 55 years, whereas an additional 35% are between the ages of 45 and 54 years (CAPCA, 2006). An initiative that promotes development of novel new approaches for managing pests, integrates management approaches into higher level IPM systems, and trains a new generation of IPM professionals should become a national priority.

References AliNiazee, M. T. (1984). Effect of two synthetic pyrethroids on the predatory mite, Typhlodromus arboreus, in the apple orchards of western Oregon. In “Acarology VI” (D. A. Griffiths and C. E. Bowman, eds.), pp. 655–658. Wiley Interscience, New York. Antle, J. M., and Park, S. K. (1986). The economic of IPM in processing tomatoes. Calif. Agric. 40(3/4), 31–32. Benbrook, C. M., Groth, E., Halloran, J. M., Hansen, M. K., and Marquartdt, S. (1996). “Pest Management at the Crossroads.” Consumers Union, Yonkers, NY. Bentley, W. J., Zalom, F. G., Barnett, W. W., and Sanderson, J. P. (1987). Population densities of Tetranychus spp. (Acari: Tetranychidae) after treatment with insecticides for Amyelois transitella (Lepidoptera: Pyralidae). J. Econ. Entomol. 80, 193–200. Blair, B. D. (1986). Dissemination of pest management information in the Midwest, USA. In “Advisory Work in Crop, Pest, and Disease

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Management” (J. Paltri and R. Ausher, eds.), pp. 231–233. SpringerVerlag, Berlin. Bolkan, H. A., and Reinert, W. R. (1994). Developing and implementing IPM strategies to assist farmers: an industry approach. Plant Dis. 78, 545–550. Brodt, S., Zalom, F., Krebill-Prather, R., Bentley, W., Pickel, C., Connell, J., Wilhoit, L., and Gibbs, M. (2005). Almond growers rely on pest control advisers for integrated pest management. Calif. Agric. 59, 242–248. Brunner, J. F. (1994). Integrated pest management in tree fruit crops. Food Rev. Int. 10, 135–157. California Association of Pest Control Advisers (CAPCA) (2006). “Pest Control Adviser Demographic Profile.” CAPCA, Sacramento, CA. Calkins, C. O., and Faust, R. J. (2003). Overview of areawide programs and the program for suppression of codling moth in the western USA directed by the United States Department of Agriculture–Agricultural Research Service. Pest Manage. Sci. 9, 601–604. Carson, R. (1962). “Silent Spring.” Houghton Mifflin, Boston. Carter, J. E. (1979, August 2). Integrated Pest Management Memorandum from the President. Available at http://www.presidency.ucsb.edu/ws/ index.php?pid32691. Cate, J., and Hinkle, M. (1993). “Integrated Pest Management: The Path of a Paradigm.” National Audubon Society, Washington, DC. Coble, H. D., and Ortman, E. A. (2004). National Road Map for Integrated Pest Management. Available at http://www.csrees.usda.gov/nea/pest/pdfs/ ipm_roadmap_5-3-04.pdf. Coble, H. D., Bonanno, A. R., McGaughey, B., Purvis, G. A., and Zalom, F. G. (1998). “Feasibility of Prescription Pesticide Use in the United States, Issue Paper 9.” Council for Agricultural Science and Technology, Ames, IA. Council on Environmental Quality (1972). “Integrated Pest Management.” Council on Environmental Quality, Washington, DC. Croft, B. A. (1990). “Arthropod Biological Control Agents and Pesticides.” Wiley, New York. Croft, B. A., and Hoyt, S. C. (1978). Considerations for the use of pyrethroid insecticides for deciduous fruit pest control in the U.S.A. Environ. Entomol. 7, 627–630. Flint, M. L., and van den Bosch, R. (1981). “Introduction to Integrated Pest Management.” Plenum, New York. Flint, H. M., Yamamoto, A. K., Parks, N. J., and Nyomura, K. (1993). Aerial concentrations of gossyplure, the sex pheromone of the pink bollworm (Lepidoptera: Gelechiidae), within and above cotton fields treated with long-lasting dispensers. Environ. Entomol. 22, 43–48. Frisbie, R. E., and McWhorter, G. M. (1986). Implementing a statewide pest management program for Texas, USA. In “Advisory Work in Crop Pest and Disease Management” (J. Palti and R. Ausher, eds.), pp. 234– 262. Springer-Verlag, Berlin. Georghiou, G. P. (1986). The magnitude of the resistance problem. In “Pesticide Resistance: Strategies and Tactics for Management.” National Academy Press, Washington, DC. Green, T. A., Gouge, D. H., Braband, L. A., Foss, C. R., and Graham, L. C. (2007). IPM STAR certification for school systems: rewarding pest management excellence in schools and childcare facilities. Am. Entomol. 53, 168–174. Gruys, P. (1982). Hits and misses. The ecological approach to pest control in orchards. Entomol. Exp. Appl. 31, 70–87. Gubler, W. D. (1991). Powdery mildew: epidemiology and control. In “Proceedings of the Nelson J. Shaulis Viticultural Symposium,” pp. 44–47. New York State Agricultural Experiment Station, Geneva, NY. Hoppin, P., Liroff, R. A., and Miller, M. M. (1996). “Reducing Reliance on Pesticides in Great Lakes Basin Agriculture.” International Policy Program, World Wildlife Fund, Washington, DC.

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Hoskins, W. M., Borden, A. D., and Michekbacher, A. E. (1939). Recommendations for a more discriminating use of insecticides. Proc. 6th Pacific Sci. Conf. 5, 119–123. Hoy, M. A., Barnett, W. W., Reil, W. O., Castro, D., Cahn, D., Hendricks, L. C., Coviello, R., and Bentley, W. J. (1982). Large scale releases of pesticide-resistant spider mite predators. Calif. Agric. 36 (1/2), 8–10. Hoy, M. A., Barnett, W. W., Hendricks, L. C., Castro, D., Cahn, D., and Bentley, W. J. (1984). Managing spider mites in almonds with pesticide-resistant predators. Calif. Agric. 38(7/8), 18–20. Hoyt, S. C. (1969). Integrated chemical control of insects and biological control of mites on apples in Washington. J. Econ. Entomol. 62, 74–86. Jacobsen, B. J. (1997). Role of plant pathology in integrated pest management. Annu. Rev. Phytopathol. 35, 373–391. Jimenez, M. J., Toscano, N. C., Flaherty, D. L., Ilic, P., Zalom, F. G., and Kido, K. (1988). Controlling tomato pinworm by mating disruption. Calif. Agric. 42(11/12), 10–12. Kogan, M. (1988). Integrated pest management theory and practice. Annu. Rev. Entomol. 49, 559–570. Kogan, M. (1998). Integrated pest management: historical perspectives and contemporary developments. Annu. Rev. Entomol. 43, 243–270. Krause, R. A., and Massie, L. B. (1975). Predictive systems: modern approach to disease control. Annu. Rev. Phytopathol. 13, 31–47. Madden, L., Pennypacker, S. P., and McNab, A. A. (1978). FAST, a forecast system for Alternaria solani on tomato. Phytopathology 68, 1354–1358. Metcalf, R. L., and Luckmann, W. H. (1982). “Introduction to Insect Pest Management.” Wiley, New York. Michelbacher, A. E. (1945). The importance of ecology in insect control. J. Econ. Entomol. 38, 129–130. Michelbacher, A. E., and Bacon, O. G. (1952). Walnut insect control in northern California. J. Econ. Entomol. 45, 1020–1027. National Research Council (1989). “Alternative Agriculture.” National Academy Press, Washington, DC. Newsom, L. D. (1980). The next rung up the integrated pest management ladder. Bull. Entomol. Soc. Am. 26, 369–374. Pickel, C., Hasey, J., Bentley, W., Olson, W. H., and Grant, J. (2002). Pheromones control oriental fruit moth and peach twig borer in cling peaches. Calif. Agric. 56, 170. Pitblado, R. E. (1992). “The Development and Implementation of TOMCAST: A Weather-Timed Fungicide Spray Program for Field Tomatoes.” Ontario Ministry of Agriculture and Food, Ontario, Canada. Prokopy, R. J. (1994). Integration in orchard pest and habitat management: a review. Agric. Ecosyst. Environ. 50, 1–10. Rajotte, E. G., Kazmierczak, R. S., Norton, G. W., Lambur, M. T., and Allen, W. A. (1987). “The National Evaluation of Extension’s Integrated Pest Management (IPM) Programs,” VCES 491-011-024. Virginia Cooperative Extension Service, Blacksburg, VA. Rice, R. E., and Kirsch, P. A. (1990). Mating disruption of oriental fruit moth in the United States. In “Behavior-Modifying Chemicals for Insect Management” (R. L. Ridgway, R. M. Silverstein, and M. N. Inscoe, eds.), pp. 193–211. Dekker, New York. Robertson, J. L., and Worner, S. P. (1990). Population toxicology: suggestions for laboratory bioassays to predict pesticide efficacy. J. Econ. Entomol. 83, 8–12. Saenz de Cabazon Irigaray, F. J., Zalom, F. G., and Thompson, P. B. (2007). Residual toxicity of acaricides to Galendromus occidentalis and Phytoseiulus persimilis reproductive potential. Biol. Contr. 40, 153–159. Sall, M. A. (1980). Epidemiology of grape powdery mildew: a model. Phytopathology 70, 338–342.

Chapter | 8  Pesticide Use Practices in Integrated Pest Management

Smith, R. F. (1974). Origins of integrated control in California, an account of the contributions of Charles W. Woodworth. Pan-Pac. Entomol. 4, 426–440. Smith, R. F., and Allen, W. W. (1954). Insect control and the balance of nature. Sci. Am. 190(6), 38–92. Sorenson, A. A. (1993). “Regional Producer Workshops: Constraints to the Adoption of Integrated Pest Management.” National Foundation for Integrated Pest Management Education, Austin, TX. Sorenson, A. A. (1994). “Proceedings of the National Integrated Pest Management Forum.” Center for Agriculture in the Environment, American Farmland Trust, De Kalb, IL. Stark, J. D., and Banken, J. A. O. (1999). Importance of population structure at the time of toxicant exposure. Ecotoxicol. Environ. Saf. 42, 282–287. Stark, J. D., Jepson, P. C., and Mayer, D. F. (1995). Limitations to use of topical toxicity data for predictions of pesticide side effects in the field. J. Econ. Entomol. 88, 1081–1088. Stern, V., Smith, R. F., van den Bosch, R. F., and Hagen, K. S. (1959). The integrated control concept. Hilgardia 29, 81–97. Stevenson, W. R. (1983). An integrated program for managing potato late blight. Plant Dis. 67, 1047–1048. United Nations Food and Agriculture Organization (UNFAO) (1967). “Report of FAO Panel of Experts on Integrated Pest Control.” UNFAO, New York. U.S. Department of Agriculture, National Agricultural Statistics Service (USDA NASS) (1998). “1997 Pest Management Practices. Special Circular 1(98).” USDA, Washington, DC. U.S. Environmental Protection Agency (1997). “Guidelines for Expedited Review of Conventional Pesticides under the Reduced-Risk Initiative and for Biological Pesticides, Pesticide Registration Notice 97-3.” U.S. Environmental Protection Agency, Washington, DC. U.S. General Accounting Office (U.S. GAO). (2001). “Agricultural Pesticides; Management Improvements Needed to Further Promote

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Integrated Pest Management”, report No. GAO-01-815. U.S. GAO, Washington, DC. Vandeman, A., Fernandez-Cornejo, J., Jans, S., and Lin, B. H. (1994). “Adoption of Integrated Pest Management in U.S. Agriculture,” Agricultural Information Bulletin 707. U.S. Department of Agriculture, Economic Research Service, Washington, DC. van den Bosch, R. (1964). Practical application of the integrated control concept in California. Proc. Intern. Congr. Entomol. 12, 595–597. van den Bosch, R., and Stern, V. M. (1962). The integration of chemical and biological control of arthropod pests. Annu. Rev. Entomol. 7, 367–386. Way, M. J. (1977). Integrated control – Practical realities. Outlook Agric. 9, 127–135. Wearing, C. H. (1988). Evaluating the IPM implementation process. Annu. Rev. Entomol. 33, 17–38. Weber, E., Gubler, W. D., and Derr, A. (1996). Powdery mildew controlled with fewer fungicide applications. Winegrowing Jan/Feb, 13–16. Wigglesworth, V. B. (1950). The science and practice of entomology. Adv. Sci. 7, 154–161. Wilson, L. T., Hoy, M. A., Zalom, F. G., and Smilanick, J. M. (1984). The within-tree distribution and clumping pattern of mites in almond orchards: comments on predator–prey interactions. Hilgardia 52(7), 1–13. Zalom, F. G. (1993). Reorganizing to facilitate the development and use of integrated pest management. Agric. Ecosystems Environ. 46, 245–256. Zalom, F. G., and Strand, J. F. (1990). Expectations for computer decision aids in IPM. AI Appl. Nat. Res. Manag. 4(1), 53–58. Zalom, F. G., Hoy, M. A., Wilson, L. T., and Barnett, W. W. (1984). Presence–absence sequential sampling for web-spinning mites in almonds. Hilgardia 52(7), 14–24. Zalom, F. G., Walsh, D., Stimmann, M. W., Pickel, C., Krueger, W., Buchner, R., and Brazzle, J. (1998). Impact of pyrethroids on beneficial mite predators. In “Proceedings of the California Plant and Soil Conference,” pp. 62–67. Agronomy Society of America, California Chapter, Sacramento, CA.

Chapter 9

Properties of Soil Fumigants and Their Fate in the Environment Husein Ajwa1, William J. Ntow1, Ruijun Qin1, and Suduan Gao2 1 2

University of California, Davis, California USDA-ARS, Parlier, California

9.1  Introduction Soil fumigants are pesticides that are used to control a wide array of soil-borne pests including nematodes, pathogens, and weeds. Soil fumigants, after application to soil, rapidly form gas via either volatilization or chemical transformation. By diffusion in the soil pores and partitioning into soil aqueous phase, the fumigant gas plays the primary role to control soil-borne pests. Soil fumigants are used intensively for pre-plant pest control in many annual crops (e.g., potatoes, tomatoes, strawberries, peppers, and carrots), nurseries (e.g., fruit trees, nut trees, and grapevine), and floriculture. Therefore, soil fumigation has become an important agricultural practice to ensure better crop yield and provide greater benefits to growers worldwide. In the United States, the majority of fumigants are applied in California and Florida (Ajwa et al., 2003). Soil fumigants currently in use are listed in Table 9.1. Most are halogenated compounds. Methyl bromide (MeBr) has been used as an effective broad-spectrum soil fumigant for several decades. About 68,400 metric tons were used worldwide in 1996, half of which was used in the United States (Ware and Whitacre, 2004). MeBr is predominantly used for pre-plant soil treatments, which accounted for 70% of the global total. Quarantine uses account for 5–8%, while 8% is used to treat perishable products, such as flowers and fruits, and 12% for nonperishable products, like nuts and timber. Approximately 6% is used for structural applications, such as for drywood termite fumigation of infested buildings (C&E News November 9, 1998). An update on the status of MeBr can be viewed at the following U.S. Environmental Protection Agency (U.S. EPA) website: http://www.epa.gov/ozone/mbr/. Methyl bromide was identified as a compound that contributes to ozone depletion in the stratosphere and an Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

international agreement (Montreal Protocol) was establi­ shed for its gradual phase-out [United Nations Environment Programmes (UNEP), 1995]. As of January 2005, MeBr was officially phased out in the United States under provisions of the U.S. Clean Air Act and the Montreal Protocol (U.S. EPA, 1994). Developing countries have until 2015 to phase out MeBr production under the Montreal Protocol (C&E News Nov. 9, 1998). Some MeBr use in the United States is under critical use exemptions (CUE) or quarantine pre-shipment (QPS) allocations, which are subject to annual regulatory approval. The phase-out of MeBr has resulted in increased use of alternative fumigants (CDPR, 2005; Trout, 2006). Some of these fumigants were used long before or jointly with MeBr, such as 1,3-dichloropropene (1,3-D) and chloropicrin (CP, trichloronitromethane). Fumigant availability is dependent on registration status. In the United States, pesticides must be registered by the U.S. EPA and sometimes by the state to be sold and distributed. In October 2007, the U.S. EPA approved the registration of methyl iodide (MeI or iodomethane) under highly restrictive provisions governing its use (U.S. EPA, 2007). Methyl iodide (MeI) is currently registered as a fumigant in many states, but not in California. U.S.-registered fumigants include primarily the following five active ingredients: MeBr, 1,3-D, methyl isothiocyanate [MITC; a primary breakdown product of metam-sodium/ potassium (methyldithiocarbamate) or Basamid (Dazomet)], CP, and MeI. Other chemicals, such as propargyl bromide and sodium azide, were evaluated as alternatives to methyl bromide (Ajwa et al., 2003) but were not considered for registration in the United States. The U.S. EPA has granted experimental use permits (EUP) for other fumigants (e.g., dimethyl disulfide or DMDS) (Howard, 2007). The permits allow designated growers on commercial farms to evaluate formulations containing unregistered active ingredients. 315

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Table 9.1  Physicochemical properties of soil fumigants Molecular formula

Molecular weight (g/mol)

Boiling point (°C)

Density (g/ml)

Water solubility (g/l)

KH Vapor pressure (kPa)

Kd or Kf (ml/g)

t1/2

Rf

Af L  25 cm

Methyl bromide

CH3Br

94.9

3.6

1.73 (0°C)

13.4(25°C)

227 (25°C)

0.04–0.10

4–52

2.37

0.59

Methyl iodide

CH3I

141.9

42.4

2.28(20°C)

14.0(25°C)

53 (25°C) 0.21 (25°C)

n.a.

5–43

n.a.

n.a.

cis-1.3-D

C3H4Cl2

111.0

104.3

1.22(20°C)

2.32(25°C)

4.5 (25°C) 0.074 (25°C)

0.5–1.5

3–17

2.81

0.04

trans-1.3-D

C3H4Cl2

111.0

112

1.22(20°C)

2.18(25°C)

3.1 (25°C) 0.043 (25°C)

0.4–0.70

3–17

2.79

0.02

Chloropicrin

Cl3CNO2

164.4

112

1.66(20°C)

1.62(25°C)

3.2 (25°C) 0.10 (20°C)

0.14–0.03

0.2–4

n.a.

n.a.

MITC

CH3NCS

73.1

118–119

1.05(24°C)

8.2 (25°C)

2.5 (20°C) 0.01 (20°C)

0.012

1–13

1.34

0.37

Dimethyl disulfide

C2H6S2

94.2

110

1.06(16°C)

4.2

2.9 (20°C) 0.05 (20°C)

0.57

1.53

n.a.

Carbon disulfide

CS2

76.1

45.5

1.26(20°C)

2.94

47 (25°C) 0.078 (10°C)

n.a.

0.90

n.a.

0.24 (20°C)

KH , Henry’s law constant (dimensionless); Kd or Kf , linearized adsorption or Freundlich coefficient; Koc, soil organic carbon sorption coefficient; t1/2 , half-life; Rf , retention factor; Af, attenuation factor; n.a., data not available.Basic chemical parameters are from “The Pesticide Manual” (Tomlin, 1994) and Ajwa et al. (2003). References for Kd , Kf , are from Ajwa (2002), Frick (1995), Gan et al. (1996), Kim et al. (2003a, b); Rf and Af are from Yates et al. (2003). Half-life data are summarized information for normal soil conditions from numerous studies discussed in the text.

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Fumigant

Chapter | 9  Properties of Soil Fumigants and Their Fate in the Environment

None of the alternatives (except MeI) effectively control as wide a range of soil-borne pests as MeBr. For example, although 1,3-D is effective against nematodes, it lacks herbicidal activity and is often formulated with chloropicrin, which is a better fungicide. MITC is effective against nematodes and a variety of weeds and fungal pathogens (Dungan and Yates, 2003). However, MITC fumigants should not be applied with 1,3-D or CP simultaneously because 1,3-D and CP transformation can be accelerated in the presence of MITC (Zheng et al., 2003). Therefore, MITC is often applied sequentially with CP or 1,3-D to better control weeds. Methyl iodide is the most promising compound that provides pest control that is equal to or better than MeBr (Becker et al., 1995; Ohr et al., 1996; Sims et al., 1995). The chemical structures of registered soil fumigants as well as some nonregistered fumigants are shown in Figure 9.1.

317

Reviews of various soil fumigants, particularly MeBr and its alternatives, are available (Ajwa et al., 2003; Dungan and Yates, 2003; Ruzo, 2006; Yates et al., 2003). This chapter provides basic and updated information on the chemical properties and environmental fate of currently used soil fumigants, environmental issues surrounding these fumigants, and strategies to minimize any negative impacts from soil fumigation.

9.2  Chemical properties, application, and major environmental issues Fumigants are volatile organic compounds, which means they are capable of transforming and producing volatile ingredients. They become gases at relatively low temperatures after they are applied to soils. Major physiochemical

H H

Methyl bromide

C

Br

H H Iodomethane

H

C

I

H Cl O Chloropicrin

Cl

C

N O

Cl

Cl 1,3-Dichloropropene

CH2Cl C

C

C

H

H

Dimethyl disulfide

Carbon disulfide Figure 9.1  Chemical structure of fumigants.

H3C

CH2Cl Trans-1,3-D

N

H3C

C

H

Cis-1,3-D

Methyl Isothiocyanate

H

Cl

C

S=C=S

C

C

S

CH3

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properties of soil fumigants are given in Table 9.1. These compounds have generally low boiling points, high vapor pressure, and low solubility that reflect their high volatility and potential to partition into gas phase. Henry’s law constant (KH) is a measure of fumigant concentration ratio in gas phase over its concentration in liquid phase at equilibrium and it can be used to evaluate the volatility of a chemical:



C K H  air Cwater

(1)

where Cair and Cwater are the concentrations of fumigant in air and water, respectively, upon establishing an equlibrium. KH is dimensionless, whereas both Cair and Cwater are measured in mg/l. Fumigants with high KH values have greater tendency to transfer from liquid to gas phase and to distribute easily over a large area. Considering the large air volume in soil, fumigant loss to the air can be very high. However, the KH values for all fumigants are less than unity, indicating that most fumigants would partition to aqueous phase. Thus, increasing soil water content would lead fumigant partitioning more in the aqueous phase and retain fumigants in the soil. Fumigant vapor pressure, solubility, and KH are all dependent on temperature. Specifically, as temperature increases, vapor pressure and KH increase while solubility decreases (Yates et al., 2003). Thus, comparison of these values is only valid under similar temperature conditions.

9.2.1  Fumigation Methods Various fumigant application techniques are used depending on the formulation type, pests to be controlled, and timing of the application (Lembright, 1990).

9.2.1.1  High-Pressure Liquid Gas Injections Methyl bromide applications are performed as liquid gas injections with nitrogen gas or pressurized air. The fumigant is injected directly into the soil via tractor-driven shanks or chisels (shank injection). The fumigant is usually applied from 15 to 60 cm below the soil surface through shank injections. Spacing of the shanks can vary from 20 to 50 cm depending on injection depth. Deeper injection depth allows further movement of fumigants through soil pores and allows wider shank spacing.

9.2.1.2  Low-Pressure Liquid Chisel Injections Applications of 1,3-D (Telone II, Telone C35) are typically made as liquid chisel injections. Chloropicrin and metam sodium (Vapam) or metam potassium (K-pam) liquids may also be applied in this manner. The liquid fumigant is

pumped by a positive displacement pump through a system of valves into hoses that lead to injection chisels (shanks). Flushing out is generally accomplished with nitrogen gas or pressurized air.

9.2.1.3  Power Mulch/Rototiller Applications These applications can be used when target pests are in the upper soil profile. The fumigation material is applied to the field surface, as either a liquid spray or a granular broadcast, and immediately incorporated to the desired depth with a power mulching device (rototiller). For granular formulation applications, the spreading equipment must be able to adequately handle the small granule size and maintain delivery to the area where the granules will be physically incorporated. Drop spreaders are generally preferable to broadcast spreaders because of the reduced potential for drift and off-target applications. Once the granular or liquid fumigant has been applied and mechanically incorporated, the soil surface is often sealed with a plastic tarpaulin or a sprinkler irrigation application (University of California, 2009).

9.2.1.4  Chemigation Applications Fumigants can also be delivered into soil via irrigation systems such as sprinklers or drip tapes (drip application). This method is called chemigation. The irrigation water acts as a vehicle for pesticide distribution and can provide a more uniform and deeper distribution of the fumigant (Ajwa and Trout, 2004). Water solubility of fumigants is generally low, although it is high enough for soil pest control. Some fumigants such as 1,3-D and CP can be applied directly with irrigation water. However, emulsified formulas (e.g., InLine containing 61% 1,3-D, 33% CP, and 6% inert ingredient) are commonly used for drip application. Drip tapes are installed either near the surface (e.g., in raised beds for strawberry production) or subsurface (15 cm or deeper) and application may last a few hours (Ajwa and Trout, 2004). Granule formulations such as dazomet are applied to the soil surface and then watered via sprinklers into the soil to generate MITC.

9.2.1.5  Current Innovative Fumigation Approaches Regulatory restrictions on the use of several soil fumigants due to growing concern over their negative environmental attributes underline the need for careful application and consideration of a combination of methods to achieve efficacy without negative impact on the environment. Although numerous approaches are currently being investigated and implemented in many countries, only a handful are surveyed and described in Section 9.6.

Chapter | 9  Properties of Soil Fumigants and Their Fate in the Environment

9.2.2  Environmental Concerns As pesticides, soil fumigants are toxic compounds that are used in large quantities ranging from 100 to 400 lb/ acre. Without control, large amounts of the fumigant can be emitted into the atmosphere due to its highly volatile nature. Many pesticide active and inert ingredients are also identified as volatile organic compounds (VOCs) (Segawa, 2008). Methyl iodide does not contribute to ozone formation. VOCs are compounds which are defined such that upon release to the atmosphere they can react with nitrogen oxides under sunlight to form harmful ground level ozone:

VOCs + NOx sunlight  → Ozone

(2)

Ozone is a key ingredient of urban smog and an air pollutant harmful to living materials. Thus, regulations (e.g., use limits and buffer zones in the United States) have been used to minimize emissions and protect public and environmental health. Stringent fumigant use regulations are being developed to reduce air emissions especially in some air-quality no attainment areas in California (e.g., CDPR, 2008; Segawa, 2008). Some of these fumigants are under re-evaluation and new or more stringent regulations are expected in the near future to increase protection for agricultural workers and bystanders (U.S. EPA, 2008). Minimizing emissions or reducing potentially negative impacts on the environment is the key to maintaining the practical use of fumigants in agriculture.

9.3  Processes and factors affecting the fate of fumigants in soil A number of simultaneous processes affect the fate of fumigants in soil, as illustrated in Figure 9.2. Fumigants applied to the soil form or transform quickly to gases

Figure 9.2  Processes affecting the fate of fumigants in soil.

319

that are subject to partitioning to soil air, water and solid phase (most importantly organic matter) through diffusion and sorption, volatilization (emission), degradation, and potential leaching. Emission loss is the major concern affecting potential exposure risks to workers and bystanders and impairment of air quality. Emission loss can also reduce fumigant efficacy when it results in reduced fumigant residence time in the soil rooting zone. The ultimate goal of soil fumigation is to achieve maximum control of pests, which requires an effective concentration and uniform distribution of fumigants in soil. Thus, containment of fumigants in soils is necessary to ensure maximum control. Volatilization and degradation of fumigants in soils through chemical reaction or microbial activities are considered the major pathways that affect the dissipation of fumigants after they are applied to soils.

9.3.1  Volatilization The high volatility of fumigants ensures a high degree of diffusion within soil that is beneficial for pest control. However, volatility results in emission losses without proper containment strategies. Figure 9.3 illustrates the emission flux of 1,3-D and CP using various application methods and surface treatments (irrigation or tarping). Without tarp or water barriers, emissions to the atmosphere can be very high. Fumigant emissions from soils are affected by soil conditions (texture, moisture, and organic matter content), weather, application methods, and surface barriers, as well as fumigant properties. In general, lower emissions are expected from soils with fine texture, high water content, high soil organic matter (SOM) content, and low temperature compared to soils with coarse texture, dry, low SOM content, and high temperature conditions. Approaches to reduce fumigant emissions include management of application methods, physical barrier, irrigation,

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amendment with chemicals or organic materials, and target area treatment. These measures are detailed in Section 9.6.

9.3.2  Degradation Fumigant degradation in soil is attributed to chemical reaction and biodegradation. Chemical reactions include hydrolysis or nucleophilic substitution of halogens on fumigants with water (hydrolysis; Eq. 3) or various functional groups (e.g., NH2, NH, SH, OH, COOH) on organic materials (methylation; Eq. 4) (Gan et al., 1994; Guo et al., 2004). For example, for MeBr or MeI, these reactions can be described as

CH3 X  H 2 O → CH3 OH  X  H

(3)

CH3 X  OM-XH → CH3 OM-H  X  H

(4) where X represents Br or I and OM-H represents neucleophilic functional groups of SOM, such as fulvic acids, that form bound residues with MeBr and 1,3-D (Xu et al., 2003).

Figure 9.3  Emission flux of (a) 1,3-dichloropropene (1,3-D) and (b) chloropicrin (CP) from Telone C35 (shank-injection) and InLine (dripirrigation) applications in a sandy loam soil with or without surface treatments. Control, bare soil; pre-irrigation, irrigate surface 12 in. of soil to its field capacity prior to fumigation; HDPE, high-density polyethylene; VIF, virtually impermeable film. Error bars are standard deviations of duplicate measurements (source: Gao et al., 2008, JEQ).

Biodegradation occurs when microorganisms are able to metabolize pesticides and utilize them as a source of energy and nutrients. The contribution of chemical and biological mechanisms to fumigant degradation largely depends on soil conditions and timing. Chemical and biological mechanisms are often determined by comparing fumigant degradation in sterilized and unsterilized soils and the results vary regarding dominant mechanisms. Chemical mechanisms accounted for 91% of the total 1,3-D degradation in mineral soils and 58% in composted steer-manureamended soils. Minimal biodegradation was reported in soils without fumigation history. Hydrolysis can play an important role shortly following fumigant application (Guo et al., 2004). Gan et al. (1998a) found that sterilization slightly increased the 1,3-D degradation in soils amended with biosolid-manure mixtures, indicating the dominance of chemical reactions in fumigant breakdown. Others, however, reported the primary role of microbial activity on 1,3-D dissipation in soils with and without organic matter (OM) amendment (Gan et al., 1998a; Ibekwe et al., 2001; Ma et al., 2001). It is reasonable to assume that fumigation also plays a role in soil disinfection so microbial degradation would not be important until the microbial population had sufficient time to recover. Degradation rate or half-life (t1/2) is often used to evaluate fumigant degradation. First-order kinetics or user-defined availability adjusted first-order kinetics models are often used to describe degradation data (Guo and Gao, 2009). Many factors including fumigant characteristics and soil conditions (temperature, SOM, texture, water content, and to a lesser degree soil pH, mineralogy, or adsorption capacity) affect degradation rate. The half-life of fumigants due to degradation in soils ranging in texture under normal conditions (i.e., without amendment or treatment) generally follows the order of MeBr (4–52 days)  MeI (14–32 days)  1,3-D (3–17 days)  MITC (1–13 days)  CP (0.2–4 days) (Table 9.1). Fine-textured soil with high clay content, SOM, and water content would lead to low half-life values for a fumigant. Increasing temperature reduced fumigant half-life by increasing chemical or biological degradation rate. A high degradation rate of fumigant in a soil may require a higher application rate to ensure good efficacy. Enhancing degradation at the soil surface (e.g., apply chemicals or OM) is only proposed to minimize emissions. Environmental factors such as weather and soil conditions affect one or more of the degradation processes, and thus the fate of fumigant in soil.

9.3.3  Adsorption Fumigants can bind to soil particles including minerals and OM. To describe fumigant adsorption in soils, an adsorption coefficient (Kd) is used for linear adsorption isotherm: Kd  Cs/Cw, where Cs is fumigant concentration on soil in mg/kg, and Cw is the concentration in aqueous phase

Chapter | 9  Properties of Soil Fumigants and Their Fate in the Environment

in mg/l. For a wide concentration range including high fumigant concentrations, nonlinear adsorption isotherms are often observed and the Freundlich equation is often used:

Cs  K f Cwn

where Kf is the Freundlich coefficient and n is a constant describing the nonlinearity of the adsorption isotherm. When n  1, Kf  Kd. The observed n ranged from 0.93 to 0.94 for 1,3-D isomer adsorption in soils (Kim et al., 2003a). Most Kd or Kf values (Table 9.1) are very low (1) suggesting most fumigants are weakly adsorbed. However, peat soils and OM-amended soils resulted in much higher Kd or Kf because OM has a much higher affinity to fumigants than minerals. Kim et al. (2003a) found that the Kf is positively related to OM content in soils:

For cis-1, 3-D, K f  0.31 [OM(%)]  0.43



For trans -, 3- D, K f  0.29[OM(%)] − 0.25



To evaluate the role of OM for fumigant adsorption, Koc is defined as an expression of adsorption capacity based on organic carbon content (foc):



K oc 

K f (or K d ) foc



Because foc is generally less than unity, Koc is often much higher than Kd or Kf. For soils, higher Kd, Kf , or Koc values indicate that fumigants are more strongly attached to the soil. Available Koc values are 18–60 for 1,3-D isomets (Kim et al., 2003a) and 22 for MeBr (Wauchope et al., 1992). Organic amendment into surface soil has been proposed to promote degradation as well as adsorption that may serve as a strategy to reduce emissions.

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factor (Af). The Rf is an index of the relative time needed for a pesticide to move past some specified depth compared to a nonadsorbing tracer and the Af is the fraction of pesticide mass that is likely to move past some specified depth including the effects of adsorption and degradation (Yates et al., 2003). Some of the Rf and Af values are given in Table 9.1. These values indicate that MeBr is highly mobile but can be more persistent in soils than 1,3-D isomers. MITC has much lower Rf values; therefore, it is less likely to achieve uniform distribution in soils than MeBr and 1,3-D isomers and may require higher concentrations to achieve an effective distribution in soil profile. Indeed, fumigant distribution in soils often varies greatly with location, soil depth, and application time, depending on fumigation method, injection depth, and shank or driptape spacing (McKenry and Thomason, 1974; Wang et al., 2004). Figure 9.4 shows typical distribution patterns of 1,3D using shank injection. Figure 9.5 provides examples of 1,3-D distribution in raised beds from shank-injected and drip-applied fumigants from a strawberry field. Fumigant in the soil-gas phase usually starts with the highest concentration at the injection depth or location and diffuses fairly quickly throughout the soil. Figure 9.4 shows that within 12–24 h, similar concentrations were achieved in locations at the injections lines and between injection lines. Figure 9.4 also shows that fumigant concentration in soil decreases dramatically over time primarily due to emission loss and degradation. A fumigant is often evaluated for its efficacy on soil pest control using a dosage response curve that compares soil pest control efficiency with fumigant exposure over time. A fumigant exposure index (C  T, where C is fumigant concentration in mg/l and T is the time in h) is used to assess efficacy and to determine the lethal dosage (e.g., LD95 and LD99.99) or a threshold value for specific pest control (McKenry and Thomason, 1976; Wang et al., 2004). The dosage ( concentration  time) exposure to soil pests at a specific location and depth can also be visualized as

9.4  Fumigant distribution in soil and efficacy assessment Adequate concentration, uniform distribution, and sufficient residence time of fumigants in soil are critical for effective pest control. Gaseous fumigants in soil pores are primarily driven by diffusion. Fumigant distribution from chemigation would be affected by the transport of water, heat, and fumigants. Various models and several mobility indices have been used to describe fumigant transport or mobility in soils under various conditions (Yates et al., 2003). These indices can be used to compare the transport potential of fumigants in soil. More detailed information regarding these indices can be found in Yates et al. (2003). Two of the parameters are retardation factor (Rf) and attenuation

Figure 9.4  1,3-Dichloropropene distribution in soil-gas phase at (a) fumigant injection (shank) line and (b) center between shank lines for a bare soil. Application rate, shank injection depth, and shank spacing were 500 kg/ha, 45 cm, and 50 cm, respectively.

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1,3-D concentration in the soil gas (mg L−1 air)

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9.5  Currently available soil fumigants

Drip Fumigation

3.0

Bed center Bed edge

2.5

9.5.1  Methyl Bromide

2.0 1.5 1.0 0.5 0.0 0

10

20

30

40

50

60

70

80

1,3-D concentration in the soil gas (mg L−1 air)

Time after application (hours)

Shank Injection

3.0

Bed center Bed edge

2.5 2.0 1.5 1.0 0.5 0.0 0

10

20

30

40

50

60

70

80

Time after application (hours) Figure 9.5  1,3-Dichloropropene concentration (mg/l air) in the soil gaseous phase of a silty loam soil after drip fumigation with InLine (top) and shank injection with Telone C35 (bottom) at 325 l/ha at 30 cm depth (source: Schneider et al., 2008).

the area under the concentration curve in Figure 9.5. The threshold value can be used to determine effective fumigant application rate depending on application methods and soil conditions such as texture, water content, and bulk density. McKenry and Thomason (1976) determined the dosage in soil where fumigant concentration in soil aqueous phase was used and converted from measured gaseous concentration based on Henry’s law constant. Other reported values (Schneider et al., 2008; Wang et al., 2004) used directly soil gaseous fumigant concentration integrated with time. To ensure good efficacy, soil fumigation must achieve the threshold value for soil pests throughout the treated areas. Thus, soil conditions, such as high soil water content, can inhibit gaseous fumigant diffusion and reduce efficacy. For fine-textured soils, the effect of soil water content on fumigant diffusion was most striking when soils had soil-water tension in excess of 50 kPa at a 30-cm depth (McKenry and Thomason, 1974).

Methyl bromide is a colorless, nonflammable, low boiling point chemical with high vapor pressure (227 kPa at 25°C) and reasonable water solubility (13.4 g/l) (Yates et al., 1996). It is often applied by shank injection as a liquid (from pressurized cylinder), which quickly vaporizes and diffuses in soil. Without tarps, as much as 21–87% of the applied MeBr can be released to the atmosphere (Yagi et al., 1993, 1995; Yates et al., 1997). The Henry’s law (KH) constant for MeBr is 0.24 at 20°C (Table 9.1). The adsorption coefficient (Kd) in different soil types is negligible but could increase to 0.20 in soils with high organic matter content (e.g., potting mix). The predominant mechanism that induces the spreading of MeBr through the soil profile is vapor diffusion (Goring, 1962; Kolbezen et al., 1974; Reible, 1994). After injection, which may involve a short period where pressure-driven flow dominates, liquid MeBr vaporizes and moves throughout the soil in response to the phase-change expansion and the initially high gradients near the injection points. Degradation of MeBr is mainly via chemical hydrolysis and methylation through a SN2 nucleophilic substitution with water and nucleophilic sites on soil OM as indicated in Eqs. 3 and 4 (Gan et al., 1994). Bacteria have also been implicated by the oxidation of MeBr (Miller et al., 1997; Oremland et al., 1994; Ou et al., 1997; Rasche et al., 1990). This reaction is catalyzed by monooxygenase (Dungan and Yates, 2003):

CH3 Br  1 / 2O2 → H 2 CO  H  Br

(5)

Rasche et al. (1990) found that two soil ammoniaoxidizing nitrifiers, Nitrosomonas europaea and Nitroso­lobus multiformis, consumed MeBr only in the presence of ammonium chloride. Inhibition of biodegradation by allylthiourea and acetylene, specific inhibitors of the ammonia monooxygenase, suggests that the enzyme catalyzed MeBr degradation. Oremland et al. (1994) showed that a methanotrophic bacterium, Methyloccus capsulatus, was also capable of co-oxidizing MeBr when incubated in the presence of methane. Methyl bromide did not support growth of the methanotroph. Miller et al. (1997), however, isolated a gram-negative aerobic bacterium that was able to utilize MeBr as a sole carbon and energy source. Detailed information about MeBr degradation can be found in Dungan and Yates (2003). For pre-plant soil fumigation, mechanized shank injection is the predominant application method (Yates et al., 2003). Tarping with either high- or low-density films (HDPE or LDPE) following fumigant injection is often used to retain fumigants for good efficacy and to reduce emissions. Deep injection (45 cm or deeper) with tarps is

Chapter | 9  Properties of Soil Fumigants and Their Fate in the Environment

required under newly adopted environmental regulations (CDPR, 2008). Shank spacing varies from 25 cm to 2 m depending on injection depth that considers movement of MeBr in the soil profile. MeBr is often applied with a small amount of CP (e.g., 98:2 MeBr/CP). Application rates range from 240 to 480 kg/ha (United Nations Environment Programmes, 1995). Lower rates can be used for tarped treatments.

9.5.2  Methyl Iodide (Iodomethane) MeI is considered as the best candidate to replace MeBr for soil fumigation because of its relatively equivalent efficacy in controlling a variety of soil-borne pests including weeds, nematodes, and fungi (Becker et al., 1995; Ohr et al., 1996; Sims et al., 1995). The main advantage of MeI over MeBr is that it degrades quickly in the troposphere via photolysis and therefore is unlikely to contribute to ozone depletion in the upper stratosphere (Gan and Yates, 1996). The estimated life of MeI in the atmosphere is 4–8 days compared with 1.5–2 years for MeBr; its estimated ozone depletion potential (ODP) is only 0.016 compared with 0.6–0.7 for MeBr (Gan and Yates, 1996). In terms of ground-level ozone formation, MeBr and MeI are not VOCs (i.e., they do not contribute to ozone formation). Methyl iodide is structurally analogous to MeBr and has many similar physical–chemical properties (Figure 9.1 and Table 9.1). In comparison to MeBr, MeI has a higher boiling point and a lower vapor pressure. As a result, distribution of MeI between the soil–water–air phases showed slower movement in the soil profile and less volatilization from the soil surface compared to MeBr (Gan and Yates, 1996). The degradation of MeI is also similar to that of MeBr. Degradation of MeI occurs mainly through chemical hydrolysis and methylation through a SN2 nucleophilic substitution with water and nucleophilic sites on SOM, respectively (Eqs. 3 and 4). The half-life of MeI from degradation in some soils in closed vials can be over 30 days at 20°C (Guo and Gao, 2009). However, the persistence of MeI in soil was found to decrease with increasing soil organic matter content (Gan and Yates, 1996; Guo and Gao, 2009). Soil texture, mineralogy, and moderate soil water content change had little influence on MeI degradation (Guo and Gao, 2009).

9.5.3  1,3-Dichloropropene 1,3-D has been considered as one of the most viable alternatives to MeBr (Noling and Becker, 1994) and has been widely used as a pre-plant control of parasitic nematodes and fungi. Commercial formulations for 1,3-D are registered under the names Telone II (975 g AI/kg, Dow Agrosciences) and D-D (Shell). Both contain nearly equal

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concentrations of the cis- and trans- isomers (Figure 9.1). Typical application rates for Telone II soil fumigant for field crop use on mineral soils range from 130 to 195 kg/ha, with 388 kg/ha being the maximum rate (Batzer et al., 1996). Products with a mixture of 1,3-D and CP (e.g., Telone C17, Telone C35) are often used. These products offer satisfactory pest control in many field tests (Schneider et al., 2008). An emulsified formulation of 1,3-D is InLine (61% 1,3-D, 33% CP, and 6% inert ingredient), which has commonly been applied through drip irrigation systems in strawberry fields in coastal areas of California. 1,3-D isomers have relatively lower vapor pressure, lower KH, higher sorption coefficients, and higher degradation rates compared to MeBr (Table 9.1). With the sufficiently high vapor pressure, 1,3-D movement in soil is dominated by gas phase diffusion but at a slower rate than MeBr. The higher sorption coefficients of 1,3-D may indicate that 1,3-D can persist in soils. Yates et al. (2003) estimated a slightly higher Rf for 1,3-D (2.8) than MeBr (2.4). In fact, the maximum depth of detectable residues in soil dissipation studies was less than 3 m. This movement has been considered to be the result of diffusion rather than leaching. Extensive studies have indicated that dissipation of 1,3-D in soils is attributed to both hydrolysis and metabolism. The hydrolysis of 1,3-D is a mixed SN1/SN2 reaction with water or OH- as nucleophiles and produces 3-chloroallyl alcohols, which can be further transformed to carboxylic acid intermediates and finally to CO2 (Batzer et al., 1996; Guo et al., 2004). The hydrolysis process can be described as

ClCH 2  CH  CHCl  H 2 O → HOCH 2  CH  CHCl  Cl−  H

or



ClCH 2  CH  CHCl  OH → HOCH 2  CH  CHCl  Cl

The hydrolysis of 1,3-D isomers had a similar half-life range of 9.8–12 days at 20°C (Guo et al., 2004; McCall, 1987; Wang et al., 2000). In soil, it is hypothesized that cis- and trans-1,3-D are initially hydrolyzed to corresponding cis- and trans-3chloroallyl alcohol (Dungan and Yates, 2003). The isomers of 3-chloroallyl alcohol are then oxidized to cis- and trans3-chloroacrylic acid, which can subsequently be degraded to succinic acid, propionic acid, and acetic acid. The aliphatic carboxylic acids can also be mineralized to CO2, H2O, and Cl–. Detailed information on the mechanisms of 1,3-D degradation in soil can be found in Dungan and Yates (2003). The degradation of 1,3-D in aerobic soils has been examined by numerous investigators (Batzer and Yoder, 1995; Batzer et al., 1996; Dungan et al., 2001; Dungan and Yates, 2003; Gan et al., 1998a; Jeffers and Wolfe, 1996; Leistra, 1970; Leistra et al., 1991; Roby and Melichar,

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1996; van Dijk, 1980) with observed half-lives ranging from 2 days (silty clay) to about 6 days on sandy loams or clay soils and 17 days in sandy soils at 20°C. The major metabolites are 3-chloroallyl alcohol, 3-chloroacrylic acid, and carbon dioxide. The degradation rates of the 1,3-D isomers were mostly reported to be similar in soils (Leistra et al., 1991; Smelt et al., 1989; van Dijk, 1980; van der Pas and Leistra, 1987), with relatively greater degradation of trans1,3-D than cis-1,3-D in enhanced soil (Dungan and Yates (2003). The half-life of 6–7 days was typically observed in sandy loam soils at 20°C and with a soil water content of 10% (Kim et al., 2003; Wang et al., 2000). Increasing soil moisture may or may not contribute to 1,3-D degradation (Gan et al., 1999; Dungan et al., 2001). Increasing application rate from 10 to 500 mg/kg resulted in the increase in half-life of 1,3-D from 5 to 14 days (Dungan et al., 2003b). Increasing soil temperature from 20 to 50°C resulted in accelerated 1,3-D degradation rates of three to seven times those in a sandy loam soil due to increased microbial metabolism and chemical reaction rates (Dungan et al., 2001; Gan et al., 1999; Ma et al., 2001). The half-life of 1,3-D was much shorter in high SOM soil compared to that in low SOM soil (1.2 vs. 5.3 days) for the same type of soil (Ashworth and Yates, 2007).

9.5.4  Chloropicrin CP fumigation has been used for many decades to control soil-borne pests. The early use and development of CP as a soil fumigant is reviewed elsewhere (Noling, 1997; Wilhelm, 1966; Wilhelm and Paulus, 1980). CP is typically applied together with MeBr as a warning agent because of its property as a lacrimator (tear producer). It is often used with 1,3-D as a fungicide to achieve broad-spectrum control. For instance, products of MeBr with 2% CP and 1,3-D with 17% or 35% CP (i.e., Telone C-17 and Telone C-35) have been formulated by Dow AgroSciences. CP and its combinations with 1,3-D or sequential application with MITC have been identified as effective replacements for MeBr in many field studies (Freitas et al., 1999; Moldenke and Thies, 1996; Porter et al., 1999; South et al., 1997; Trout and Ajwa, 1999). CP has also been recognized as an effective agent for controlling replant disorder problems in orchards (Browne et al., 2003; Trout et al., 2003). Replant disorder is a generic lack of vigorous growth when the same crop is planted into a field year after year. Potentially contributing factors to replant disease include populations of soil-borne pathogens and nematodes, as well as unidentified factors. CP treatment sometimes performed better than other fumigants for replant disease. CP effectively prevented almond replant disorder when 0.5–2.0 lb was used per tree site (Greg et al., 2003). CP is a clear, colorless, nonflammable liquid with moderate vapor pressure and boiling point (Figure 9.1,

Hayes’ Handbook of Pesticide Toxicology

Table 9.1). The physicochemical parameters for CP transport in soil have been reported by Wilhelm et al. (1996). Compared to 1,3-D, CP has lower pressure but higher KH and a much shorter half-life. As a result, its dissipation in soils is often faster than 1,3-D, although initial distributions in soil were similar (Gao and Trout, 2007; Gao et al., 2008b). Subsequent lower emissions were also observed consistently when an equivalent amount of fumigant was applied (Qin et al., 2008a). Degradation of CP is much faster than MeBr, MeI, 1,3-D, and MITC as reflected in its much shorter half-life (Table 9.1). The half-life of CP was in the range of 1.5–5.8 days in sandy loam soil at 20°C (Gan et al., 2000b; Wang et al., 2000). CP degradation in soils is attributed to both hydrolysis and biodegradation. However, the hydrolysis of CP is much slower than 1,3-D with a half-life of 83.7 days at 20°C (Wang et al., 2000). In three soil types (Arlington sandy loam, Carsitas loamy sand, and Waukegan silt loam), the half-life of CP is 1.5, 4.3, and 0.2 days, respectively, under aerobic conditions (Dungan and Yates, 2003). Sterilization of these soils increased the half-life of CP to 6.3, 13.9, and 2.7 days, respectively, indicating that biodegadation plays a major role in CP degradation. Microbial degradation was estimated to account for 60–90% of the overall CP degradation in soils (Gan et al., 2000b). Under anaerobic/aquatic conditions, CP is converted to nitromethane within hours. The major metabolic pathway occurs through three successive reductive dehalogenations to nitromethane: Cl3 CNO2 → Cl2 CHNO2 → ClCH 2 NO2 → CH3 NO2 A small portion (about 4%) of the chloropicrin was also converted to CO2 (Dungan and Yates, 2003). In both aerobic and anaerobic environments significant binding of radiocarbon to soil fulvic and humic fractions was observed. The final breakdown products of chloropicrin are carbon dioxide, nitrate, and chloride. When applied with other fumigants, competitive degradation between CP and 1,3-D has been reported in amended and unamended soils (Desaeger et al., 2004; Zheng et al., 2003). Repeated fumigation generated accelerated CP degradation compared to the soil without fumigation history. The degradation of CP was highly affected by soil temperature but not soil moisture (Gan et al., 2000b). In three types of soils (sandy loam, loamy sand, and silt loam), when the soil temperature increased from 20 to 50°C, CP degradation rate increased by 7.5, 11.0, and 7.0 times, respectively (Gan et al., 2000b).

9.5.5  Methyl Isothiocyanate MITC is the primary breakdown product from metam sodium or potassium salts (Figure 9.1) and is considered as the active ingredient for soil-borne pest control. MITC

Chapter | 9  Properties of Soil Fumigants and Their Fate in the Environment

is a broad-spectrum pesticide with activity against plant pathogenic nematodes, weeds, oomycota, and a variety of plant pathogenic fungi (Duniway, 2002). Metam sodium has been distributed under a variety of tradenames since the 1950s (e.g., Vapam HL, 42% metam sodium, Amvac Chemical Corp., Newport Beach, CA). Metam sodium is typically applied as a 37 wt% solution in water (Draper and Wakeham, 1993). MITC can also be generated in soil by applying the granular product dazomet (tradename Basamid; BASF Corp., Mount Olive, NJ). Both metam sodium and dazomet convert efficiently to MITC in moist soils with half-lives in the order of hours to days, depending on ambient conditions. These products are not used as stand-alone fumigants but have been found to be useful when applied sequentially with other fumigants such as 1,3-D. MITC is unstable and decomposes to methylamine in water, probably via thiocarbamic acid. Faster hydrolysis rates are obtained at lower pH levels. MITC is much less susceptible to acid catalysis in water than its oxo-analog but will react with a great variety of nucleophiles. Because of its high vapor pressure (19 mmHg at 20°C; Table 9.1), it is important to understand its photolysis in the vapor phase. Photolysis of MITC in the gas phase proceeds with a half-life of 10 h using a xenon arc lamp and nearly 1 day under sunlight. This rapid rate stands in contrast to that in aqueous solutions, where the reaction is 20 times slower. Multiple products are observed, including methyl isocyanide, sulfur dioxide, hydrogen sulfide, N-methylformamide, methylamine, and carbonyl sulfide. Methyl isocyanide, in turn, degrades to methyl isocyanate (Ruzo, 2006). MITC is weakly sorbed and because of its volatility and water solubility, it can partition into both vapor and water phases. Thus, it comes into contact with soils through leaching and diffusion. MITC exhibits only moderate diffusive mobility when compared with other soil fumigants. Particularly, inclusion of organic amendments into the soil surface enhances degradation. For instance, Dungan and Yates (2003) reported that in soil, the degradation rate of MITC was about six times higher when amended with 5% composted chicken manure. The addition of compost or manure adds more organic matter to the soil which can develop a new microbial population with enhanced degradation capacity for MITC (Zhang et al., 2005). The degradation of MITC is also influenced by soil temperature and moisture content. In a sandy loam soil, the degradation rate of MITC was about three times higher at 40°C than at 20°C (Dungan et al., 2002). Changes in the soil moisture content below saturation had little influence on the degradation rate of MITC in this soil, but in contrast to 1,3-D, degradation of MITC was 2.6 times slower at a soil moisture content of 16% than at 1.8% in a loamy sand soil (Gan et al., 1999). Thus, MITC degradation decreased with increasing soil moisture content but increased with increasing soil temperature, so its effect would be magnified in the hot, dry surface layer of soil (Gan et al., 1999).

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Degradation of MITC is attributed to both biological and chemical mechanisms (Dungan and Yates, 2003). At 20°C, microbial degradation accounted for as much as 50–80% of the total degradation. The accelerated degradation of other carbamate pesticides by adapted microorganisms has also been reported (Felsot et al., 1981; Rahman et al., 1979). Repeated applications of MITC to soils also appear to enhance degradation as a result of increased populations of adapted microorganisms (Smelt et al., 1989).

9.5.6  Dimethyl Disulfide Dimethyl disulfide (DMDS) is a new pre-plant soil fumigant being developed by Arkema on a worldwide basis for the treatment of nematodes, weeds, and soil-borne plant pathogens. DMDS is a ubiquitous natural product, common in the global sulfur cycle, and is detected as a metabolite in numerous biological processes. DMDS is not only malodorous but also very toxic for all organisms. It exerts a complex mode of action through mitochondria dysfunction and activation of ATP-sensitive potassium channels and greatly inhibits the cytochrome oxidase (Auger et al., 2002, cited in Fritsch, 2005). The product is being evaluated in the United States and other countries under EUP (Howard, 2007). The properties of DMDS are similar to the properties of MITC (Table 9.1). However, the behavior of DMDS in soil is not well documented. Recent research indicated that the half-life of DMDS in soil is two to three times greater than the half-life of MITC.

9.5.7  Sodium Tetrathiocarbonate Sodium tetrathiocarbonate (STTC) is formulated as Enzone. It was first manufactured by Unocal Corp., Chemical Division, and currently is the product of Arysta Lifescince, Inc. Enzone is a deep amber-colored, nonexplosive liquid formulation of STTC that breaks down in the soil to carbon disulfide (CS2) gas, the active moiety. Enzone is registered as both a pre-plant and a post-plant fumigant but is used primarily post-plant in established orchards or vineyards in California. Additional registrations outside the United States include its use for vegetables, raspberries, and strawberries. STTC is not as volatile as other fumigants (e.g., MeBr, CP, and 1,3-D) and does not move as easily with the soil air. It moves through the soil profile to the target pests more efficiently with soil moisture (Rf value 0.9) (Adaskaveg, 1999; http://mbao. org/1999airc/97philli.pdf, accessed October 11, 2008). Based on manufacturer’s recommendations, STTC can be used at high concentrations as a pre-plant fumigant or at low concentrations as a post-plant treatment possibly without causing phytotoxicity to growing plants at application sites.

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The dissociation of STTC occurs by dilution or hydrolysis as described below:

Na 2 CS4  2 H 2 O → CS2  H 2 S  S  2 NaOH

The hydrolysis reaction is very fast and occurs within 2 h at water pH values of less than 9. Our research indicated that the generation of CS2 in soil after the application of STTC is instantaneous and that a complete generation of CS2 occurs within 1 h after application to sandy loam and clay loam soils. Therefore, the distribution of CS2 depends on the application method that controls the distribution of the water-soluble formulation (Enzone).

9.6  Strategies to minimize emissions Fumigants are subject to fast volatilization and over half of fumigants applied to soil can be lost through emission without proper containment strategies. Emission loss results in poor efficacy on soil-borne pests and the release of contaminants that degrades air quality. More stringent environmental regulations are expected to minimize emissions from soil fumigation. This section summarizes the current information about several techniques that can minimize fumigant emissions while maintaining adequate efficacy.

9.6.1  Application Methods For a given application method, deeper injection would always lead to lower emissions. Increasing shank injection depth from 25 to 60 cm showed 20% or more emission reductions of MeBr in bare soils (Yates et al., 2002). For bare soil fumigation, a general consensus is that emissions from drip, especially subsurface drip application, are always lower than broadcast shank injections (Gao et al., 2008a; Wang et al., 2008). This is because increasing soil water content decreases air pore volume for diffusion and increases the amount of fumigant partitioning in the aqueous phase. In addition, with drip fumigation, there are no shank traces serving as volatilization channels. Moreover, with shank injection, substantially high soil water content could reduce fumigant distribution in soils in contrast to drip application where fumigant moves with the applied water (Ajwa and Trout, 2004).

9.6.2  Plastic Films Plastic tarp (mulch) or film has been primarily used to contain fumigants in soil and to control fumigant emissions. The effectiveness of tarping on emission reductions depends largely on the chemical characteristics and tarp permeability. A mass transfer coefficient (h, cm/h) is used to describe tarp permeability; its measurement is described in Papiernik et al. (2001). The higher the h values, the higher the tendency for fumigants to pass through the film.

Tarping with polyethylene (PE) film was traditionally used for MeBr but found ineffective to control 1,3-D emissions especially in relatively dry soils (Wang et al., 1999; Papiernik and Yates, 2002). However, a HDPE (high-density PE film) tarp applied over an irrigated soil profile resulted in lower 1,3-D emissions that were attributed to both higher soil water content and water condensation under the film (Gao and Trout, 2007). Tarped treatment in an irrigated field, however, could improve soil pest control due to elevated soil temperature under the tarp (solarization effect) (Shrestha et al., 2006). In recent years, low-permeability materials, typically called virtually impermeable film (VIF), showed great potential to reduce emissions in early laboratory or small plot tests (Wang et al., 1997a). VIF is generally a multi­ layered film composed of barrier polymers such as nylon or ethyl vinyl alcohol (i.e., EVOH) sandwiched between PE polymer layers (Villahoz et al., 2008). A number of studies have shown that VIF can retain higher fumigant concentrations than HDPE, thus reducing emissions while improving efficacy especially for weed control (Hanson et al., 2008; Noling, 2002; Wang et al., 1997b). More recent data confirm that this type of film can effectively reduce emissions (Ajwa, 2008; Qin et al., 2008a; Yates et al., 2008b). The effectiveness of VIF on emission reductions in large field applications may be reduced as the result of damage done to the film during field installation. Tarp permeability did increase after field installation, but its permeability was still substantially lower than PE films (Ajwa, 2008).

9.6.3  Irrigation or Water Treatment Proper irrigation management can minimize emissions. Intermittent water applications with sprinklers following fumigation (water seals) can reduce MITC, 1,3-D, and CP emissions in the field. (Gao and Trout, 2007; Sullivan et al., 2004; Yates et al., 2008a). The effect is more pronounced on reducing emission peak flux (up to an 80% reduction) following fumigant application than total or cumulative emission loss because water is only applied during the first few days (Gao et al., 2008b). Nevertheless, reducing the peak flux is important because it reduces the potential exposure risk to workers and bystanders during fumigation. Buffer zones are determined based on the peak emission flux. However, high water content in the surface (0–15 cm) soil could reduce surface pest control such as nematodes and weeds (Hanson et al., 2008).

9.6.4  Chemical Amendment Soil amendments with chemicals [e.g., ammonium or potassium thiosulfate (ATS or KTS), thiourea, or polysulfides] are extremely effective to reduce emissions. These chemicals can react with fumigants such as MeBr, 1,3-D,

Chapter | 9  Properties of Soil Fumigants and Their Fate in the Environment

CP, and methyl iodide to form nonvolatile compounds by dehalogenation (Gan et al., 1998b; Wang et al., 2000). The practicality of using these chemicals on a large field scale is inconclusive at this time due to cost factors and potentially undesired soil/fumigant/thiosulfate reactions that leave odors lasting for a long time in the field (Gao et al., 2008c). Zheng et al. (2007) indicated that the smell may have derived from sulfur by-products from the transformation of thiosulfate and fumigants in soil. In some urgent cases such as spills, these chemical treatments would be effective.

9.6.5  Organic Amendment Soil amendment with organic materials such as composted manure has shown effectiveness in degrading fumigants and also reducing emissions in laboratory and field studies. Because of the strong adsorption of fumigants into organic matter (Xu et al., 2003), soil with high SOM content was reported to give lower emissions (Ashworth and Yates, 2007). Field data regarding the efficiency of organic amendment to reduce fumigant emissions are inconclusive. One study reported amendments were sometimes effective (Yates et al., 2008b) and another study reported effectiveness when up to 10 tons/acre of composted dairy manure was applied (Gao et al., 2008c). The uncertainty may relate to the reaction kinetics between fumigants with OM, the quantity as well as the quality of organics.

9.6.6  Target Area Treatment Fumigation to target areas such as tree rows or tree sites may be applicable for some orchards where pre-plant disease is the major concern in preventing establishment of healthy crops. Shank application of fumigants in row-strip (shank-strip) or drip-application of fumigant in tree site (drip-spot) have been proposed and tested in fields for efficacy of alternative fumigants (Browne, 2008). These target area treatments reduce emissions by reducing total treatment acreage to less than 50% (shank-strip) or 10% (drip-spot) of total field area. Reducing treatment areas in the field is not recommended if a field has nematode populations.

9.6.7  Mass Balance of Fumigants   Applied to Soil The fate of fumigants after soil application has been evaluated in various studies. This information can be better presented from laboratory controlled studies. It is commonly found that over a period of 1 or 2 weeks, about half or greater of applied fumigants can be easily lost through volatilization and the remainder through degradation (Gao and Trout, 2007; McDonald et al., 2008, 2009). Residual

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fumigants measured in the soil were minor (3%). The amount of fumigant lost to degradation can increase substantially due to water treatment and/or surface sealing or amendment with chemicals or organic materials. The amount of fumigant that is subject to degradation often increases as emission losses decrease.

Conclusion Soil fumigation will continue to be critical to sustain agricultural production as the world’s population grows. Increased public awareness of environmental issues surrounding fumigants is leading to more stringent regulations toward the safe use of these volatile compounds and minimal release into the environment. Management strategies must be developed for various agronomic systems to maximize fumigation effects on soil-borne pest control with minimal input and to minimize any negative environmental impact.

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thiosulfate in furrows on chloropicrin emissions from drip fumigated fields. Chemosphere 72, 558–563. Qin, R., Gao, S., and Ajwa, H. (2008b). Low permeable tarps reduce emissions from drip-applied inline in a strawberry field trial. In “Proceedings of the Annual International Research Conference on MeBr Alternatives and Emission Reductions, Nov. 11–14, 2008. Orlando, FL,” pp. 114 (1–4). http://www.mbao.org/2008/Proceedings/mbrpro08.html. Rahman, A., Atkinson, C., Douglas, J. A., and Sinclair, D. P. (1979). Eradicane causes problems. N.Z. J. Agric. 139, 47–49. Rasche, M., Hyman, M. R., and Arp, D. J. (1990). Biodegradation of halogenated hydrocarbon fumigants by nitrifying bacteria. Appl. Environ. Microbiol. 56, 2568–2571. Reible, D. D. (1994). Loss of methyl bromide to the atmosphere during soil fumigation. J. Hazard. Mat. 37, 431–444. Roby, D. M., and Melichar, M. W. (1996). 1,3-Dichloropropene regulatory issues. In “Fumigants: Environmental Fate, Exposure and Analysis,” ACS Symposium Series 652, pp. 25–31. American Chemical Society, Washington, DC. Ruzo, L. O. (2006). Review: physical, chemical and environmental properties of selected chemical alternatives for the pre-plant use of methyl bromide as soil fumigant. Pest Manag. Sci. 62, 99–113. Schneider, S. M., Ajwa, H. A., Trout, T. J., and Gao, S. (2008). Nematode control from shank- and drip-applied fumigant alternatives to methyl bromide. HortScience 43, 1826–1832. Segawa, R. (2008). Volatile organic compound emissions from pesticides. California Department of Pesticide Regulation. http://www.cdpr. ca.gov/docs/dept/prec/2008/voc_allowance.pdf. Shrestha, A., Gao, S., and Trout, T. J. (2006). Surface water applications for reducing emissions from Telone C35: their effect on weed populations. In “Proceedings of the Annual International Research Conference on Methyl Bromide Alternatives and Emissions Reductions, Orlando, FL (6–9 November 2006),” p. 116(1–4). http:// www.mbao.org/2006/06Proceedings/mbrpro06.html. Sims, J. J., Grech, N. M., Ole Becker, J., McGiffen, M. Jr., and Ohr, H. D. (1995). Methyl iodide: a potential alternative to methyl bromide. In “Proceedings of the Annual International Research Conference on Methyl Bromide Alternatives and Emissions Reductions, San Diego (6–8 November 1995),” p. 46. Outreach, Fresno, CA. Smelt, J. H., Crum, S. J. H., and Teunissen, W. (1989). Accelerated transformation of the fumigant methyl isothiocyanate in soil after repeated application of metam sodium. Environ. Sci. Health B24, 437–455. South, D. B., Carey, W. A., and Enebak, S. A. (1997). Chloropicrin as a soil fumigant in forest nurseries. For. Chron. 73, 489–494. Sullivan, D. A., Holdsworth, M. T., and Hlinka, D. J. (2004). Control of offgassing rates of methyl isothiocyanate from the application of metamsodium by chemigation and shank injection. Atmos. Environ. 38, 2457–2470. Tomlin, C. (1994). “The Pesticide Manual.” British Crop Protection Council and the Royal Society of Chemistry, United Kingdom. Trout, T. (2006). Fumigant use in California – response to the phase-out. In “Proceedings of the Annual International Research Conference on Methyl Bromide Alternatives and Emissions Reductions, orlando, FL (6–9 November 2006),” p. 18(1–6). http://www.mbao. org/2006/06Proceedings/018TroutTmb-fumuse-06.pdf. Trout, T., and Ajwa, H. (1999). Strawberry response to fumigants applied by drip irrigation systems. In “Proceedings of the Annual International Research Conference on Methyl Bromide Alternatives and Emissions Reductions, San Diego (1–4 November 1999),” p. 10:1–3. Agricultural Research Consulting, Fresno, CA.

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Trout, T., Schneider, S., Ajwa, H., and Gartung, J. (2003). Fumigation and fallowing effects on replant problems in California peach. In “Proceedings of the Annual International Research Conference on Methyl Bromide Alternatives and Emissions Reductions,” p. 55. United Nations Environment Programme (UNEP) (1995). “The Montreal Protocol on Substances that Deplete the Ozone Layer: 1994 Report of the Methyl Bromide Technical Option Committee.” UNEP, Nairobi, Kenya. University of California (2009) “Field Fumigation,” UC IPM’s Pesticide Applicator Compendium series, Vol. 9. Statewide IPM Program, Agriculture and Natural Resources, University of California. U.S. EPA (2007). Extension of conditional registration of iodomethane (methyl iodide). http://www.epa.gov/pesticides/factsheets/iodomethane_fs.htm. U.S. EPA (2008). Implementation of risk mitigation measures for soil fumigant pesticides. http://www.epa.gov/oppsrrd1/reregistration/soil_ fumigants. Van den Berg, F., and Ross, A. H. (1994). Tuinstra LGMTh and Leistra M. Measured and computed concentrations of 1,3-dichloropropene and methyl isothiocyanate in air in a region with intensive use of soil fumigants. Water Air Soil Pollut. 78, 247–264. Van der Pas, L. J. T., and Leistra, M. (1987). Movement and transformation of 1,3-dichloropropene in the soil of flower-bulb fields. Arch. Environ. Contam. Toxicol. 16, 417–422. Van Dijk, H. (1980). Dissipation rates in soil of 1,2-dichloropropene and 1,3- and 2,3-dichloropropenes. Pestic. Sci. 11, 625–632. Villahoz, M. D., Garza, F., Barrows, P., and Sanjurjo, M. (2008). TIF (Totally Impermeable Film): an innovative film for mulch, broadcast fumigation, and greenhouses in agriculture. In “Proceedings of the Annual International Research Conference on Methyl Bromide Alternatives and Emissions Reductions, Orlando, FL (11–14 November 2008),” p. 37(1–2). http://www.mbao.org/2008/Proceedings/mbrpro08. html. Wang, D., Yates, S. R., Ernst, F. F., Gan, J., Gao, F., and Becker, J. O. (1997a). Methyl bromide emission reduction with field management practices. Environ. Sci. Technol. 31, 3017–3322. Wang, D., Yates, S. R., Ernst, F. F., Gan, J., and Jury, W. A. (1997b). Reducing methyl bromide emission with a high barrier plastic film and reduced dosage. Environ. Sci. Technol. 31, 3686–3691. Wang, D., Yates, S. R., Gan, J., and Knuteson, J. A. (1999). Atmospheric volatilization of methyl bromide, 1,3-dichloropropene, and propargyl bromide through two plastic films: transfer coefficient and temperature effect. Atmos. Environ. 33, 401–407. Wang, Q., Gan, J., Papiernik, S. K., and Yates, S. R. (2000). Transformation and detoxification of halogenated fumigants by ammonium thiosulfate. Environ. Sci. Technol. 34, 3717–3721. Wang, D., He, J. M., and Knuteson, J. A. (2004). Concentration–time exposure index for modeling soil fumigation under various management scenarios. J. Environ. Qual. 33, 685–694. Wang, D., Tharayil, N., Qin, R., Gao, S., and Hanson, B. (2008). Reducing 1,3-dichloropropene and chloropicrin emissions with subsurface drip and virtually impermeable film. In “Proceedings of the Annual International Research Conference on Methyl Bromide Alternatives and Emissions Reductions, Nov. 11–14, 2008. Orlando, FL,” pp. 33(1–3). http://www. mbao.org/2008/Proceedings/mbrpro08.html.

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Ware, G. W., and Whitacre, D. M. (2004). “An Introduction to Insecticides,” 4th ed. MeisterPro Information Resources, Willoughby, OH. http://ipmworld.umn.edu/chapters/ware.htm. Wauchope, R. D., Buttler, T. M., Hornsby, A. G., Augustijn Beckers, P. W. M., and Burt, J. P. (1992). The SCS/ARS/CES pesticide properties database for environmental decision-making. In “Reviews of Environmental Contamination and Toxicology” (G. W. Ware, ed.), vol. 123. Springer-Verlag, New York. Wilhelm, S. (1966). Chemical treatments and inoculum potential of soil. Annu. Rev. Phytopathol. 4, 53–78. Wilhelm, S., and Paulus, A. O. (1980). How soil fumigation benefits the California strawberry industry. Plant Dis. 64, 264–270. Wilhelm, S. N., Shepler, K., Lawrence, L. J., and Lee, H. (1996). Environmental fate of chloropicrin. In “Fumigants: Environmental Fate, Exposure, and Analysis,” ACS Symposium Series 652, pp. 79–93. Xu, J. M., Gan, J., Papiernik, S. K., Becker, J. O., and Yates, S. R. (2003). Incorporation of fumigants into soil organic matter. Environ. Sci. Technol. 37, 1288–1291. Yagi, K. J., Williams, J., and Wang, N. Y. (1993). Cicerone, agricultural soil fumigation as a source of atmospheric methyl bromide. Natl. Acad. Sci. USA 90, 8420–8423. Yagi, K. J., Williams, J., Wang, N. Y., and Cicerone, R. J. (1995). Atmospheric methyl bromide (CH3Br) from agricultural soil fumigations. Science 267, 1979–1981. Yates, S. R., Gan, J. Y., Ernst, F. F., Mutziger, A., and Yates, M. V. (1996). Methyl bromide emissions from a covered field. I. Experimental conditions and degradation in soil. J. Environ. Qual. 25, 184–192. Yates, S. R., Wang, D., Ernst, F. F., and Gan, J. (1997). Methyl bromide emissions from agricultural fields: bare-soil deep injection. Environ. Sci. Technol. 31, 1136–1143. Yates, S. R., Gan, J., and Papiernik, S. K. (2003). Environmental fate of methyl bromide as a soil fumigant. Rev. Environ. Contam. Toxicol. 177, 45–122. Yates, S. R., Knuteson, J., Ernst, F. F., Zheng, W., and Wang, Q. (2008a). The effect of sequential surface irrigations on field-scale emissions of 1,3-dichloropropene. Environ. Sci. Technol. 42, 8753–8758. Yates, S. R., Knuteson, J., Ernst, F. F., Zheng W., and Wang, Q. (2008b). Reducing field-scale emissions of 1,3-d with composted municipal green-waste. In “Proceedings of the Annual International Research Conference on Methyl Bromide Alternatives and Emissions Reductions, Orlando, FL (11–14 November 2008),” p. 32 (1–3). http://www.mbao.org/2008/Proceedings/mbrpro08.html. Yates, S. R., Gan, J., Papiernik, S. K., Dungan, R., and Wang, D. (2002). Reducing fumigant emissions after soil application. Phytopathology 92, 1344–1348. Zhang, Y., Spokas, K., and Wang, D. (2005). Degradation of methyl isothiocyanate and chloropicrin in forest nursery soils. J. Environ. Qual. 34, 1566–1572. Zheng, W., Gan, J., Papiernik, S. K., and Yates, S. R. (2007). Identification of volatile/semivolatile products derived from chemical remediation of cis1,3-dichloropropene by thiosulfate. Environ. Sci. Technol. 41, 6454–6459. Zheng, W., Papiernik, S. K., Guo, M., and Yates, S. R. (2003). Competitive degradation between the fumigants chloropicrin and 1,3-dichloropropene in unamended and amended soils. J. Environ. Qual. 32, 1735–1742.

Section II

Toxicity and Safety Evaluation

(c) 2011 Elsevier Inc. All Rights Reserved.

Toxicity and Safety Evaluation of Pesticides Lindsay Hanson1 and Leonard Ritter2 1 2

Ottawa, Ontario, Canada School of Environmental Sciences, University of Guelph, Guelph, Ontario, Canada

Pesticide is a general term for a wide variety of products designed to control and manage pests. The term pests, at least in the context of a statutory or regulatory definition, extends to any unwanted or undesirable species. Common examples of pesticides and pests include herbicides to control weeds, insecticides to control insects, fungicides to control certain types of plant diseases, insect repellents, rodenticides to control rats, mice, gophers and other rodents, algicides to control algae in swimming pools, antifouling agents to control organisms that attach to boat hulls, and preservatives to control the decay of wood and other material. A pesticide may be a chemical or biological (e.g., bacteria and viruses used as pest control products) control agent. Pesticides differ from many other environmental substances of concern in that they enter the environment through intentional use for specified purposes. Ironically, it is the same biological effects that make pest control products valuable to society that may also result in unwanted effects that may pose risks to human and environmental health. Pesticide use in general, and some specific pesticide uses (landscape, for example), have emerged as the focus of one of the major public policy debates of our time. The use of chemical pesticides is certainly not, however, new. Stephenson and Solomon (2007) note that the use of chemical wastes to control roadside weeds was widely practiced by the Romans more than 2000 years ago. These authors also note that the development of inorganic chemicals as herbicides was under way in the 1800s and others, most notably fungicides to control plant diseases, followed. The dawn of the chemical era for pest control likely tracks its origins to the introduction in the 1930s of dinitrophenol, the first synthetic organic chemical for the control of weeds, insects, and plant diseases (Stephenson and Solomon, 2007). Rapid Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

development followed in the 1940s and 1950s, with the synthesis of the insecticide dichlorodiphenyltrichloroethane (DDT), the herbicide 2,4-dichlorophenoxyacetic acid (2,4-D) and the fungicide captan. Public concern over pesticide use followed soon thereafter. Indeed, in 1962 Rachel Carson, an accomplished naturalist, published her landmark book Silent Spring in which she highlighted important adverse environmental impacts attributable to the widespread use of the persistent organochlorine insecticides (Carson, 1962), most notably DDT, which, for many, was the metaphor for indiscriminate pesticide use. Interestingly, in 1948, only 14 years before Carson reported important adverse environmental effects associated with the indiscriminate use of DDT, the Swiss chemist Paul Müller had been awarded the Nobel Prize in Physiology or Medicine for his synthesis of DDT, which would be a critical turning point in the global fight against typhus and malaria. More than 40 years after Carson identified important environmental concerns related to the use of the persistent organochlorine insecticides, there is now renewed international interest in the reintroduction of DDT, under very carefully supervised conditions, for the control of the malaria vector (The National Academies Press, 2004). Web technology has made scientific information readily available to the general public with the click of a mouse. While this availability of information can generally be viewed as facilitating transparency and is generally positive, it is often difficult for the lay reader to place this information in the proper context. This is apparent in the daily reports of “Pesticide use linked to …,” often a serious health hazard. The general public more often asks why pesticides are being used and is concerned that exposures through diet or occupational or bystander scenarios are contributing to health problems. There is a general public perception that all 333

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chemicals are dangerous, in particular the group known as pesticides. Robust scientific data are the basis for evidencebased health hazard assessment of the safety of pesticides and a cornerstone of regulatory programs. Pesticides, by their nature, have inherent hazards. They have been developed to control pests such as undesirable vegetation, insects, and fungi. Recognition of the need to carefully evaluate the potential risks to human health from pesticide use, when considered together with the potential benefit, is why countries around the world have developed rigorous pesticide regulatory programs. Canada and the United States, for example, use similar science-based approaches, entrenched in legislation, for human health risk assessment dealing with pesticides. This framework for the risk assessment of pesticides is well accepted internationally (International Programme on Chemical Safety, 1999) and consists of four key activities: hazard identification, hazard characterization, exposure assessment, and risk characterization. Toxicological evaluation involves identifying possible human health effects related to pesticide exposures and establishing levels of human exposure that would not result in adverse effects. As a starting point, a robust data set consisting of an extensive battery of toxicity studies, conducted primarily in laboratory animals, is required to identify and characterize the hazard potential posed by pesticides. These studies are typically carried out on a variety of mammalian species (rats, mice, rabbits, dogs). Hazard identification involves understanding the inherent toxicological properties of a chemical substance. This understanding is gained through the conduct of toxicity studies that will address both the duration of exposure (acute, short-term, or chronic) and the different routes of exposure (oral, dermal, and inhalation). Various endpoints of toxicity (reproductive toxicity, developmental toxicity, genotoxicity, chronic toxicity and carcinogenicity, neurotoxicity, immunotoxicity, etc.) will be assessed to fully identify the hazards posed by chemical pesticides (Health Canada, 2008). Manufacturers of pesticides in all of the Organization of Economic Co-operation and Development (OECD) countries are required to demonstrate the safety of their products, to both human health and the environment, and are responsible to carry out the requisite studies. These are conducted according to internationally accepted Test Guidelines under Good Laboratory Practices (GLP). Generally, the test substance used in toxicity testing will be the quality grade of the active ingredient that is produced during typical manufacturing processes, often identified as the “technical” grade of the compound. A number of studies with the actual end-use formulated product (marketed pesticide), which includes the solvents, adjuvants, and carriers, are also required for purposes of developing hazard label statements for each pesticide. It is essential that all toxicology studies identify the test material used in each study.

Hayes’ Handbook of Pesticide Toxicology

I.1  Health hazard evaluation—role in the assessment of pesticide risks to humans Hazard characterization involves defining the relationship between the dose of a chemical administered to or received by the test species and the qualitative and quantitative response to that chemical (Health Canada, 2008). It is generally assumed that there is a dose level below which the chemical will not elicit a response, that is, there is a threshold for the response. “Most responses elicited by a substance, including acute toxicity, chronic toxicity, neurotoxicity, irritation, developmental toxicity, and reproductive toxicity, are considered threshold in nature. Endpoints that have been observed to lack a threshold response (e.g., genotoxicity, carcinogenicity) are assumed to result in an increase in risk at any level of exposure and hence are subject to different risk assessment methodologies” (Health Canada, 2008). The experimental dose level at which no adverse effects are detected in a given study is deemed the no observed adverse effect level (NOAEL). The lowest dose level in a study that elicits an adverse effect is referred to as the lowest observed adverse effect level (LOAEL). An adverse effect is commonly defined as “a change in morphology, physiology, growth, development or lifespan of an organism which results in impairment of functional capacity or impairment of capacity to compensate for additional stress or increase in susceptibility to the harmful effects of other environmental influences” (International Programme on Chemical Safety, 1994). Determination of a true adverse effect is not always straightforward; “expert judgment is required to separate those effects that merely reflect the ability of an organism to adapt to a biological or chemical insult from a true adverse effect”. (Health Canada, 2008) “The evaluation of a mammalian toxicological database for a specific pesticide will yield numerous NOAELs for different toxicological endpoints. The selection of the most appropriate study, endpoint, and NOAEL for human health risk assessment takes into consideration which human subpopulations may be exposed, the route of exposure, and the anticipated duration and/or frequency of exposure”. (Health Canada, 2008) Although a comprehensive scientific database is available for most pesticides, one cannot prove scientifically that something is safe with absolute certainty. The general public increasingly demands assurances of safety with respect to chemical exposure. A rigorous pesticide regulatory system begins with sound scientific data.

I.������� 2  Toxicokinetic studies Toxicokinetic studies provide data on the absorption, distribution, metabolism, and excretion (ADME) of the chemical pesticide. In general terms, how does it enter the body,

Toxicity and Safety Evaluation of Pesticides

where does it go, and what happens to it? These studies will also provide information on other parameters of interest, including differences between small and large doses, and single versus multiple exposures. “This information is valuable in interpreting toxic effects, or lack thereof, and may assist in the extrapolation of animal toxicity data to humans” (Health Canada, 2005). Understanding the toxicokinetics of the pesticide may also enable more appropriate selection of doses and routes of administration used in many of the laboratory studies. Generally speaking, the use pattern and physical properties of the product, as well as toxicokinetic considerations, will assist in determining the appropriate route of exposure and duration of study.

I.������� 3  Acute toxicity studies Acute toxicity studies on active ingredients and end-use formulated pesticide products are necessary to determine the potential hazards from acute exposures. These studies are typically characterized by a high dose in a short time frame. Acute data are used for the development of appropriate precautionary statements and hazard symbols for pesticide product labels. Acute studies identify relative acute toxicities by different routes of exposure as well as the potential to produce irritation and sensitization (Health Canada, 2005).

I.������� 4  Short-term studies Short-term or subchronic studies provide information on the toxicity profile of the pesticide through daily repeated exposure often over a period of weeks or months depending on the animal species. Guidelines for short-term studies set the dosing period for a duration lasting up to 10% of the animal’s life span. This is often defined as 90 days in rats and mice, or 1 year in dogs. The data obtained from these short-term studies are useful in determining possible cumulative or delayed toxicity, reversibility and persistence of the adverse effect, and variability in species sensitivity. These studies also provide guidance for selecting dose levels for long-term studies.

I.������� 5  Long-term studies “Long-term daily repeated exposure studies are generally designed to investigate the chronic toxicity and oncogenic potential of the pesticide when administered to test animals over the major portion of their life span” (Health Canada, 2005); by convention, chronic toxicity and oncogenicity studies must include exposure periods of at least 90% of the anticipated life span of the test animal—interpreted to be 24 months duration in rats and 18 months in mice. Ideally, the data thus generated should identify dose—response relationships and, in the case of nononcogenic effects, a clear

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demonstration of a dose that is not associated with any adverse effect (the NOAEL) and possible effects of cumulative toxicity as well as permit assessment of the potential for neoplastic development (Health Canada, 2005).

I.������� 6  Reproduction studies “These studies provide information on the potential of the pesticide to influence the reproductive performance and function of the male and female parental animals, through assessment of effects on gonadal function, estrus cycles, mating behavior, conception, parturition, lactation, and weaning. Observation of progeny from conception through lactation and weaning may enable the detection of possible adverse effects on survival, viability, development, and behavior. These studies have a pivotal role in determining the potential sensitivity of the young animal” (Health Canada, 2005).

I.������� 7  Developmental toxicity studies “These studies, referred to in the past as teratogenicity studies, permit assessment of the potential of the pesticide to induce adverse effects on the developing embryo and fetus when administered to the pregnant female test animal during critical periods of organogenesis. Studies are generally conducted in a rodent and a nonrodent species. The teratogenic potential of the pesticide may be measured by the increased incidence or induction of congenital malformations. These studies also have a pivotal role in determining the potential increased sensitivity of the young animal” (Health Canada, 2005).

I.������� 8  Genotoxicity/mutagenicity studies Tests for genetic damage are designed to assess both gene mutations and chromosomal changes as well as the competency of DNA repair mechanisms. Genotoxicity studies can also be very helpful in determining and understanding carcinogenic potential (Health Canada, 2005).

I.������� 9  Neurotoxicity and developmental neurotoxicity studies “The neurotoxic potential of the pesticide may be assessed on the basis of behavior, neurophysiology, neurochemistry, and neuropathology. Neurotoxicity screening tests may be incorporated into several of the standard protocols for acute toxicity as well as short- and long-term repeated exposure toxicity studies. This may be accomplished through expanded histopathological examination of the brain, spinal

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cord, and peripheral nervous system, a functional observational battery of tests for general behavior and neurology, as well as autonomic and sensory assessment. Appropriate tests may also be incorporated into the standard protocol for reproduction studies for the purpose of assessing the neurotoxic potential of the pesticide in the progeny” (Health Canada, 2005). Developmental neurotoxicity (DNT) studies may also be required for specific pesticides where enhanced sensitivity to the effects of the chemical has been observed in young animals. Further testing may be appropriate for pesticides known or suspected to be neurotoxicants.

I.�������� 10  Immunotoxicity “The potential of the pesticide to affect the immune system may be discerned from hematology, blood chemistry, organ weights, and histopathology, routinely investigated in shortterm and chronic toxicity studies” (Health Canada, 2005). Specific aspects of the immune response, or elucidation of immunomodulation mechanisms, may be investigated through additional assays to help predict a chemically induced functional effect on the immune system. “These assays may be considered to further investigate lymphocyte subsets, humoral antibody-mediated immunity, as well as cell-mediated and nonspecific immunity” (Health Canada, 2005).

I.�������� 11  Endocrine disruptor potential Endocrine disruptor potential (such as interference with the production of sex hormones) is evaluated in the course of examining the information from reproduction, developmental, and short- and long-term toxicity studies. If the results of these studies indicate the need for further information regarding interference with normal endocrine function, additional testing may be required.

I.�������� 12  Mechanism of action “Ancillary studies designed to elucidate specific mechanisms of action in the test animal may be key in interpreting the toxicological properties of the pesticide” (Health Canada, 2005). Such information may permit a more appropriate assessment of the relevance of the animal studies, and potential adverse health effects identified therein, to an understanding of health hazards (Health Canada, 2005).

Conclusion The importance of a robust data set when regulating a chemical substance such as a pesticide cannot be overstated. The general public, with instant access to emerging science from around the globe, must be confident in their regulatory environment. Regulatory bodies must be keenly aware of evolving science and the questions raised by those emerging issues. They must continue to recognize and adapt to the requirements of comprehensive studies from industry to address these questions. Industry must be prepared to meet these demands if they look to address public concerns. Risk communication in an open and transparent environment will be key to building public confidence. All users of pesticides must be made aware of the importance of following label directions and heeding precautionary statements. A complete and comprehensive set of toxicological studies will provide the cornerstone for building that trust.

References Carson, R. (1962). “Silent Spring,”. Houghton Mifflin, Boston, MA. International Programme on Chemical Safety. (1994). Environmental Health Criteria 170. Assessing Human Health Risks of Chemicals: Derivation of Guidance Values for Health-Based Exposure Limits. Geneva, Switzerland, World Health Organization, International Programme on Chemical Safety. www.inchem.org/documents/ehc/ ehc/ehc170.htm International Programme on Chemical Safety. (1999). Environmental Health Criteria 210. Principles for theAssessment of Risks to Human Health from Exposure to Chemicals. Geneva, Switzerland, World Health Organization, International Programme on Chemical Safety. www.inchem.org/documents/ehc/ehc/ehc210.htm Health Canada (2005). “Pest Management Regulatory Agency Regulatory Directive DIR2005-01: Guidelines for Developing a Toxicological Database for Chemical Pest Control Products”. Health Canada Ottawa. Health Canada (2008). “Pest Management Regulatory Agency Science Policy Note SPN2008-01. The Application of Uncertainty Factors and the Pest Control Products Act Factor in the Human Health Risk Assessment of Pesticides”. Health Canada, Ottawa. Stephenson, G. R. and Solomon, K. R. (2007). “Pesticides in the Environment,”. Canadian Network of Toxicology Centres Press, Guelph, Canada. The National Academies Press (2004). Saving Lives, Buying Time: Economics of Malaria Drugs in an Age of Resistance. Board on Global Health, The National Academies Press.

Chapter 10

Risk Assessment for Acute Exposure to Pesticides* Roger Cochran California Department of Pesticide Regulation, Sacramento, California

10.1  Introduction Regulatory agencies have tended to focus their assessments of pesticide risk on the potential for toxicological effects to arise from repetitive, long-term usage of the chemicals [Barnes and Dourson, 1988; World Health Organization (WHO), 1978]. This emphasis on developing reference doses (RfDs) for the potential effects of chronic exposure to pesticides may have gained impetus from public concerns about cancer or the impact of pesticides on the environment [Carson, 1962; National Research Council (NRC), 1987]. Consequently, much of the toxicological database required under the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA) examines the effects of repetitive, subchronic or chronic dosing by the oral route. From a public health perspective, however, the recorded human illnesses attributed to acute exposures to pesticides may be of greater significance than those connected with potential chronic exposures (Mehler et al., 1992). Risk assessment, as defined by the NRC (1983), consists of five components: (1) hazard identification encompasses examination of the toxic effects of the chemical; (2) dose–response assessment evaluates the dose level of the chemical necessary to cause manifestation of toxic effects; (3) exposure assessment estimates the amount of the chemical that people are likely to absorb; (4) risk characterization predicts the likelihood that people, exposed to the chemical to the degree estimated, will become ill; and (5) risk appraisal examines the strengths and weaknesses of the estimates of the various toxicological and exposure parameters and expresses the degree of confidence in the *

The opinions expressed in this chapter represent the views of the author and do not necessarily reflect the views and policies of the Department of Pesticide Regulation. The mention of trade names or commercial products does not constitute an endorsement or recommendation for use.

Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

projected risks. The term acute exposure, here, refers to human encounters with pesticides in the course of 1 day or less. A pesticide, as defined by the U.S. Environmental Protection Agency (U.S. EPA), is any chemical, or mixture of chemicals, intended to be used in preventing, destroying, repelling, or mitigating any pest (Federal Register, 1998).

10.2  Toxicological data The first two of the five components needed for risk assessment require an extensive knowledge of the toxicological effects of a chemical. Under FIFRA, the toxicological database for a pesticide is defined by the guideline requirements (U.S. EPA, 1984). This database includes acute lethality studies (oral, dermal, and inhalation); subchronic toxicity studies (90-day oral, inhalation, and dermal toxicity; 21/28-day dermal toxicity, developmental toxicity; reproductive toxicity); chronic toxicity studies (1-year non-rodent toxicity, oncogenicity, and combined chronic toxicity/oncogenicity); and neurotoxicity studies (neurotoxicity screening battery, 90-day neurotoxicity, developmental neurotoxicity) (Federal Register, 1998). Only a few of these study types contain data that can be used to explore the toxicological effects from a single day’s (acute) exposure to a pesticide. Acute lethality studies, for example, use a range of single doses to elicit toxic effects. However, these studies are designed to set toxicity categories for labeling information (U.S. EPA, 1998a). Virtually all of the older acute lethality studies, regardless of the route of exposure (oral, dermal, or inhalation), generally do not have data on nonlethal, systemic effects that occurred at less than lethal dosages. The single-dose, neurotoxicity screening battery is currently being required only for those pesticides designed to be neurotoxins (e.g., organophosphates, carbamates, fiproles, and pyrethroids) (U.S. EPA, 1998b). Consequently, data 337

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from this test, which includes components of histopathology, tissue and blood chemistry, as well as clinical signs and performance testing, are not available for most pesticides. Thus, data on acute effects for most pesticides have to be teased out of repetitive dosing studies. Subchronic, reproductive, and chronic toxicity studies may have data concerning clinical signs that appear within 1–2 days at the beginning of the studies. All other data on potential systemic toxicity in these study types are obtained at the end of the study period and cannot be attributed to acute toxicity. Developmental toxicity studies provide an exception. Because developmental toxicity may be manifested as the result of a single dose (Ogata et al., 1984; Schardein, 1985; U.S. EPA, 1991), it is assumed, in the absence of data to the contrary, that the observed developmental effects are elicited from a single dose. This assumption may or may not be valid. Nonetheless, developmental toxicity studies (see Chapter 16) tend to be a major source of critical no-observed-effect levels (NOELs) for conducting risk assessments on potential acute exposures to nonneurotoxic pesticides. Although a developmental endpoint for exposure to toxins is only relevant in women of childbearing age, the assumption that all other population subgroups are as sensitive results in margins of safety (MOSs) that protect the health of these other subgroups for other endpoints that may occur at higher dosages. The MOS is defined as the ratio of the critical NOEL to the estimated exposure. Published research studies may also provide sufficient data for dose–response assessment. These studies, however, tend to be designed to clarify the mechanism of action of a specific type of pesticide toxicity. Nonetheless, the peer-reviewed reports sometimes describe a range of concentrations used to elicit an effect from a single dose. Such a study may provide the basis for a regulatory NOEL, particularly in the case of experiments with human subjects. The main drawbacks to these published studies are (1) the lack of individual animal data because they are typically not reported or archived and, thus, (2) the need to rely on the author’s interpretation of the results. Under FIFRA, pharmacokinetic data are sought to obtain information on how a pesticide is absorbed, distributed, biotransformed, and excreted, as well as to aid in understanding the mechanism of toxicity (U.S. EPA, 1998c). Information may also be obtained about potential tissue-specific accumulation and induction of biotransformation. Most of the pharmacokinetic data are derived from studies using the oral route of exposure. Dermal pharmacokinetic studies tend to consider solely dermal penetration and/or absorption. Pharmacokinetic studies on the inhalation of pesticides are comparatively rare, seemingly limited to fumigants. Pharmacokinetic data can have a profound effect on the dose–response assessment for a pesticide. The estimated absorbed dose of a pesticide necessary to cause toxic effects may be modified downward if there is evidence of

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less than 100% absorption through the route used in the dose–response assessment. Information regarding bioavailability via the oral route is useful, as many pesticides and their metabolites are excreted in variable amounts in the feces. Estimations of absorbed dosages from inhalation toxicity studies rely on default assumptions concerning breathing rates, tidal volumes, and chemical retention and absorption to estimate absorbed dosages (Raabe, 1986, 1988; Zielhuis and van der Kreek, 1979). Such estimates, when derived from whole-body inhalation studies, can be confounded by the fact that rats exposed to dusts or chemical vapors via whole body absorb five to eight times more material than rats exposed via nose only (Blair et al., 1974; Hext, 1991; Iwasaki et al., 1988; Jaskot and Costa, 1994; Landry et al., 1986; Langard and Nordhagen, 1980; Tyl et al., 1995; Wolff et al., 1982). The additional absorption noted in whole-body inhalation exposure studies appears to be due to an unquantifiable oral component, possibly from grooming behavior (Cochran et al., 1997). Even nose-only inhalation toxicity studies may have a significant oral component due to grooming activity (Hext, 1991).

10.3  Exposure data The second, and equally important, half of the risk assessment equation is the estimate of human exposure. The chief source of exposure to pesticides through the oral route is from the diet (Cochran et al., 1995). There can also be an oral contribution from hand-to-mouth activity in adults and children or pica in children (Binder et al., 1986; Calabrese and Stanek, 1992; Calabrese et al., 1989, 1991; Carlisle, 1992; Clausing et al., 1987; U.S. EPA, 1996). Pica in children, however, appears to be highly unusual, as only a single instance of intentional imbibing of dirt was reported out of more than 200 children whose soil ingestion from hand-to-mouth activity was documented in the preceding publications. Currently, dietary exposures are estimated by most governmental agencies through a process that combines data on dietary consumption with data on pesticide residues measured on food (Cochran et al., 1995; FAO/WHO, 1988, 1997). Dietary consumption data are generally derived from government surveys (Cochran et al., 1995; FAO/WHO, 1997; Trichopoulou, 1994; USDA, 1989–1991). Data for potential pesticide residues associated with U.S. EPA or European Union (EU) labelapproved direct food uses, as well as information about possible secondary residues in animal tissues, are also necessary for estimating human dietary exposures. These data are derived from governmental monitoring programs [California Department of Pesticide Regulation (CDPR), 1997; FAO/WHO, 1999; USDA, 1996]. However, dietary exposure to pesticides is only a fraction of the total human exposure experience.

Chapter | 10  Risk Assessment for Acute Exposure to Pesticides

Much of human occupational (persons engaged in the process of pesticide application) or nonoccupational (other than dietary) exposure to pesticides results from the handling of pesticides or other activity patterns that place people in contact with the pesticides. In general, most of the occupational and nonoccupational exposure to pesticides is through the dermal and/or inhalation routes (U.S. EPA, 1992b; Ross et al., 1992; Wolfe, 1976). Exposure estimates for these scenarios are based on environmental monitoring, passive dosimetry, or biological monitoring of individuals involved in the active handling of pesticides or engaged in activities in areas treated with those pesticides (Bonasall, 1985; Lavy and Mattice, 1986). Environmental monitoring involves measurements of pesticide concentrations in the ambient air and on surfaces. The translation of measured air concentrations into an estimated absorbed dose for humans requires assumptions on respiratory frequency, volume, and retention/absorption of the pesticide (Raabe, 1986, 1988; Frank, 2008; U.S. EPA, 1996; Zielhuis and van der Kreek, 1979). Estimation of human dermal exposure from surface concentrations of pesticides in the environment relies on the precision of various generic transfer factors (U.S. EPA, 1996). Passive dosimetry gauges air concentrations in the breathing zone and measures dermal concentrations of pesticides through the use of hand washes, dermal patches, and/or articles of clothing (Wolfe, 1976). The same assumptions for inhalation are used with air concentrations of pesticides measured in the breathing zone as were used for those detected in the ambient air. Concentrations of pesticides extracted from monitoring patches attached to the skin are assumed to be representative of chemical concentrations over a specified body surface area (Wolfe, 1976). A single ­dermal absorption value, based on submitted, chemicalspecific studies (a default of 100% has been used if specific data were not available), serves as the basis for ­estimating the absorbed dose (U.S. EPA, 1992a). It is known that the percentage of pesticide absorbed through the skin varies inversely with the concentration of the chemical (Wester and Maibach, 1976). Also, in vivo human dermal absorption is variable depending on the portion of the body to which it is applied (Feldmann and Maibach, 1974; Maibach et al., 1971; Wester and Maibach, 1985). However, at the present time, there are no scientific models available that examine the effect of multiple concentrations of pesticides on the skin, separated spatially and/or chronologically, on the absorbed daily dosage (Wester and Maibach, 1993). Biological monitoring provides an estimate of the aggregate exposure to a pesticide from all routes. Unfortunately, very few biomonitoring studies have been conducted for more than a handful of pesticides. Chemical-specific, human stoichiometric data are essential to the process of estimating absorbed dosages from excreted pesticide

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metabolites. Consequently, the principal limiting factor seems to be the lack of human pharmacokinetic data on most pesticides. Chemical-specific information is preferred for exposure data from either environmental monitoring or passive dosimetry. Surrogate exposure data (from pesticides with similar chemical and physical properties, as well as similar preparation and application practices) and generic databases, such as the Pesticide Handlers Exposure Database (PHED, 1995), are used as substitutes. The use of surrogate exposure data increases the level of uncertainty in exposure estimates. Differences in volatility between the chemical under consideration and surrogate chemicals may affect air concentrations in an unquantifiable manner. Likewise, differences in chemical properties could affect transfer factors, clothing penetration, and dermal adsorption. Differences in application rates cause assumptions to be made on the relationship between the amount of chemical handled and the amount of exposure through all routes. The principal difficulty associated with the use of PHED to estimate exposure data is that the data subsets, which are combined by the program to form work categories, are not homogeneous (van Hemmen, 1992). For example, one source of variability is that each of those studies has a different minimum detection level for the analytical method. It should be noted that the detection of dermal exposure to the body regions is not standardized. Some studies observe exposure to only selected body regions, such as the hands, arms, and face, with other body regions considered 100% protected from exposure by work clothing. Other studies have more extensive dermal measurements. Consequently, the subsets derived from the database for dermal exposure have different numbers of observations for each of the body regions. Finally, the PHED database is predicated on the relationship between the amount of pesticide handled and the degree of occupational exposure. Yet, for example, within the data set used to estimate exposures for groundboom applications without the presence of a cab, there is no correlation between the amount of pesticides being used and the amount of dermal or inhalation exposures that workers receive. The net effect of this lack of correlation between exposure and the amount of chemical used is an inability to predict, with accuracy, what exposures any worker will receive in a given work category. As explained in EPA’s policy for use of PHED data (U.S. EPA, 1999), once the data for a given exposure scenario have been selected, the data are normalized (i.e., divided by) by the amount of pesticide handled resulting in standard unit exposures (milligrams of exposure per pound of active ingredient handled). Following normalization, the data are statistically summarized. The distribution of exposure values for each body part (i.e., chest or upper arm) is categorized as normal, lognormal, or “other” (i.e., neither normal nor lognormal). A central tendency value is then selected from the distribution of the exposure values

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for each body part. These values are the arithmetic mean for normal distributions, the geometric mean for lognormal distributions, and the median for all “other” distributions. Once selected, the central tendency values for each body part are composited into a “best fit” exposure value representing the entire body. In other words, EPA uses various central tendency estimates (often the geometric mean or median, as PHED data rarely follow a normal distribution), while CDPR believes the arithmetic mean is the appropriate statistic regardless of the sample distribution (Powell, 2003). Second, for acute exposure estimates CDPR uses a 95th percentile upper bound estimate (Frank, 2007), while EPA uses a central tendency estimate for all exposure durations. Third, CDPR calculates upper 90% confidence limits for both upper bound and mean exposures, while EPA does not. Finally, there are some instances, particularly in the residential setting, in which exposures must be estimated, but no data are available. In these instances, algorithms for exposure have been developed by EPA (U.S. EPA, 1998d, 2001a, 2004). These algorithms are considered by EPA to be upper bounds for the exposure estimates.

10.4  Examples Thus far, we have examined the theoretical nature of the toxicological and exposure databases used in generating a pesticide risk assessment. How the data fit together can best be explored through critical examination of some examples of risk assessments for acute exposure to pesticides. The examples that follow were taken, for the purposes of comparison, from both risk characterization documents (RCDs) conducted for the CDPR and the concomitant Registration Eligibility Decision Documents (REDs) from EPA.

10.4.1  Methyl Parathion Methyl parathion, CAS # 298–00–0, is the common name for O,O-dimethyl O-(4-nitrophenyl) phosphorothioate. Methyl parathion is an organophosphate insecticide that acts through inhibition of acetylcholinesterase (AChE) activity. It can be used on alfalfa, almonds, barley, beans, cabbage, canola, corn, cotton, hops, oats, onions, pecans, potatoes, rice, rye, sugar beets, sunflowers, walnuts, and wheat to control insect predation.

10.4.1.1  Toxicology The lowest oral LD50 in rats was 2.9 mg/kg for males and 3.6 mg/kg for females (WHO, 1984). The dermal LD50 values in male and female rats were 41 and 46 mg/kg, respectively (WHO, 1984).

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An executive order issued in 2001 precluded EPA from considering human toxicity data in its risk assessments (U. S. EPA, 2007a,b). Consequently, EPA gauged short-term (1–30 days) occupational and other nondietary exposures with a lowest-observed-effect level (LOEL) of 0.3 mg/kgday [reduced red blood cell (RBC) and brain AChE] from a 28-day dermal toxicity study in rats (Beyrouty, 2001; U.S. EPA, 2006a). An uncertainty factor of 3 was used to extrapolate from a LOEL to a NOEL. In the study, at 0.3 mg/kg-day, a statistically significant decrease in brain AChE activity was seen in both males (cerebral cortex 21%, striatum 18%, hippocampus 19%) and females (cerebral cortex 17%, striatum 22%) on day 29 but not on day 5. No reduced RBC AChE was seen in the males at either time point. Significantly reduced RBC AChE was seen in females at 5 (20%) and 29 (28%) days. No significant reduction in plasma cholinesterase was seen at any time point in either gender at 0.3 mg/kg-day. The effects at 0.3 mg/kg-day seem equivocal compared to the consistent effects on plasma cholinesterase, RBC AChE, and brain AChE seen at 1.0 mg/kg. It should be noted that dermal absorption of methyl parathion in rats was nearly complete, based on the excretion of over 90% of the total 14C-labeled compound applied to the skin (Abu-Qare et al., 2000; AbuQare and Abou-Donia, 2000; Sved, 2001). Consequently, one would expect the NOELs from 28-day rat oral toxicity studies (0.2 mg/kg-day; Daly and Rinehart, 1980a; Kumar and Desiraju, 1992) to be no different from the NOEL in the rat dermal toxicity study. Thus, the EPA endpoint, in the absence of considering human data, should have been 0.2 mg/kg/day from the two 28-day oral toxicity studies. The endpoint CDPR used to gauge acute exposures came from an acute neurotoxicity study in rats (Koshlukova and Reed, 2003; Minnema, 1994). The single bolus dose LOEL was 7.5 mg/kg-day for clinical signs, adverse effects in the functional observation battery (FOB), and significantly reduced plasma cholinesterase, RBC cholinesterase, and brain AChE activity, with a NOEL of 0.025 mg/kg-day. It should be noted that the difference in doses was 300-fold. This allows the possibility that the “true NOEL” may be somewhat higher than 0.025 mg/kg-day. Two human toxicity studies were also reported in the CDPR RCD (Rider et al., 1969, 1970, 1971; Rodnitzky et al., 1978). In the former study, five male adults received methyl parathion for up to 33 days in daily doses up to 0.27 mg/kg-day. There were no clinical signs or symptoms and no statistically significant depression of plasma or RBC cholinesterase activity. In the latter study, two men received a dose of 0.057 mg/kg-day without any clinical signs or significant decrease in plasma cholinesterase activity. The NOEL (0.27 mg/kg-day) in humans was consistent with the subchronic oral NOELs for statistically significant reduced brain AChE activities in rats (NOEL��  ������������������������������� 0.2 mg/kg-day; LOEL  1.9 mg/kg-day) and dogs (NOEL  1 mg/kg-day; LOEL  3 mg/kg-day), coupled with concomitant reductions in plasma and RBC AChE

Chapter | 10  Risk Assessment for Acute Exposure to Pesticides

activities. Even though the human studies were not used for endpoints, at the very least no safety factor for interspecies extrapolation should have been used by either agency. Clearly, humans were not 10 times more sensitive than laboratory animals to the toxic effects of a bolus dose of methyl parathion.

10.4.1.2  Occupational Exposure The principal route of exposure for pesticide applicators using methyl parathion was through the skin. Consequently, the rate of dermal absorption is critical in estimating absorbed doses. There were no in vivo, human dermal absorption studies for methyl parathion. An in vitro dermal absorption study with human skin indicated that human dermal absorption ranged from 1.4 to 9% (Sartorelli et al., 1997). However, the study had a host of technical problems, and in vitro studies are not currently used by CDPR for establishing the rate of dermal absorption. An in vivo rat dermal absorption study indicated 96% dermal absorption of methyl parathion (Sved, 2001). This study was used as the basis for the establishment of the EPA dermal absorption rate (U.S. EPA, 2006a). However, additional data suggest that this absorption rate may not be accurate for humans. The dermal absorption of ethyl parathion, chemically very similar to methyl parathion, was 95% in rats (Shah et al., 1987). However, in vivo dermal absorption on the forearm of humans was shown to be approximately 10% (Feldmann and Maibach, 1974). One would expect a similar relationship between the percentage dermal absorption in rats and humans for methyl parathion. This value (10%), though, could not be used by CDPR as the basis for human dermal absorption of methyl parathion because acetone was used in the study as the application vehicle. Acetone disrupts the integrity of the skin surface and alters (typically increases) the rate of dermal absorption (Zendzian, 1994). Also, other data indicated that in vivo human dermal absorption of ethyl parathion, though less than 95%, is variable (ranging from 4 to 64%) depending on the portion of the body to which it was applied (Feldmann and Maibach, 1974; Maibach et al., 1971; Wester and Maibach, 1985), although the body regions with highest penetration are also typically the regions receiving the lowest dosage (Ross et al., 2001). Again, because the application vehicle was acetone, these dermal absorption rates are likely overestimates. There is another approach to estimating the rate of human dermal absorption of methyl parathion. The method, described in a published paper, came from analysis of data derived from studies of the human dermal absorption of 47 radiolabeled compounds, dissolved in acetone, applied to the ventral forearm (Durkin et al., 1995). Durkin et al. (1995) had found that there was a

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correlation between molecular weight and the human dermal absorption rate. This correlation, though, was good only for compounds with log of octanol-water partition coefficient (Kow)  1.85. Methyl parathion has a molecular weight of 263.2 and log Kow  2.8. Using Durkin’s formula (log absorption rate  0.005 MW  2.1), the dermal absorption of methyl parathion would be 6.1%. This value is in the range of the estimated dermal absorption of human skin from the in vitro studies (Sartorelli et al., 1997). The Worker Health and Safety policy, however, is to assume a default of 50% dermal absorption if there are no satisfactory in vivo dermal absorption studies (Donahue, 1996). PHED data can be used to estimate handler exposures to methyl parathion for pilots and mixer/loaders for ground applications and air-blast applicators. PHED, though useful, has limitations that prevent the use of distributional statistics on exposure estimates. For example, PHED incorporates exposure data from many studies, each with a different minimum detection level for the analytical method used to detect residues in the sampling media. Moreover, as the detection of dermal exposure to the body regions was not standardized, some studies observed exposure to only selected body parts. Consequently, the subsets derived from the database for dermal exposure may have different numbers of observations for each body part, a fact that complicates interpretation of values taken from PHED. However, in the absence of chemical-specific data, PHED provided the only data available for estimating certain handler exposures to methyl parathion. A number of uncertainties are built into PHED which can generally cause exposure estimates to be overstated. Part of this comes from the fact that approximately 70% of the inside patch data used in PHED are nondetectable values. The default assumption of using half the level of quantification to estimate exposure for those nondetects may overestimate exposure. Typically, data in PHED monitor only a small fraction of the workday. The data are then linearly extrapolated for the rest of the workday. The net effect may be an overstatement of exposure (Franklin et al., 1981; Spencer et al., 1979). The maximum amount of acres treated is always utilized by CDPR in the calculations, although mechanical problems and bad weather may tend to combine to reduce the acreage treated in actual practice. Finally, it is always assumed that there is a linear relationship between the amount of pesticide handled and the amount of dermal exposure. However, as the exact nature of the relationship has not been demonstrated, this adds to the uncertainty of the exposure estimate. EPA also uses PHED to estimate handler exposure; however, EPA approaches PHED data somewhat differently than CDPR. The differences between acute exposure estimates calculated for this chapter according to current CDPR and EPA policies are summarized in Table 10.1 for an example scenario, airblast applicator.

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Table 10.1  Comparison of Ground Mixer/Loader Exposure to Methyl Parathion Estimated from Surrogate Data (PHED) by CDPR and EPA According to Their Respective Policies Exposure estimate

Exposure rate (g AI/lb handled)a

Acute ADD (g/kg-day)b

From PHED, according to CDPR policy

406c

265c

From PHED, according to EPA policyd

102

58

a

Total exposure rate, dermal plus inhalation. Acute absorbed daily dosage (ADD) estimates assumed a maximum application rate of 1 lb a.i./acre, maximum rate on field-grown ornamentals, and an 8-h workday. Amount treated was assumed to be 100 acres treated/day (Haskell, 1998), except for EPA estimate, which assumed 80 acres treated/day (U.S. EPA, 2001b). Dermal absorption was assumed to be 50%, inhalation absorption was assumed to be 100%, and body weight was assumed to be 77 kg (CDPR) or 70 kg (U.S. EPA, 1996). c Upper-bound rate defined by the CDPR policy. d Average rate used for EPA exposure estimates calculated from values obtained from EPA Policies (U.S. EPA, 1999). Data taken from M/L, Open System, Liquids (With Gloves) in PHED. b

Finally, it should be noted that the use of surrogate pesticide data (like PHED) and default assumptions necessarily incorporate additional uncertainties in the exposure estimate. Such an inference can be drawn from comparing the estimated exposures of mixer/loaders (Table 10.2). In this instance, chemical-specific data from a biomonitoring study indicated that mixer/loaders involved in aerial applications (that handle five times the amount of methyl parathion as mixer/loaders engaged in ground applications) have exposures that were one-tenth of the mixer/loaders for ground applications. As biomonitoring data are considered superior to passive dosimetry data (Woollen, 1993), appropriate chemical-specific exposure data would be preferred. Biological monitoring studies provided data on how much of a methyl parathion metabolite, 4-paranitrophenol, appeared in the urine of some workers (mixer/loaders for aerial applications, ground boom applicators, cotton scouts, corn harvesters, and walnut harvesters) (Belcher, 2001a,b; Rotondaro, 2002; Willard, 2000a,b, 2001). In order to convert this measured parameter into an estimate of the pesticide’s absorbed dose, there must be a human pharmacokinetic study which demonstrates how much of an ingested dose of methyl parathion is excreted as 4paranitrophenol within a given period of time. The total amount of paranitrophenol excreted in the human pharmacokinetic study averaged 27% in a 24-h period, with no detectable metabolite in the urine after 24 h (Morgan et al., 1977). Yet, despite isolation in hotel rooms, nearly all of the workers in the biomonitoring studies had

paranitrophenol in their urine during the 48 h before the studies began. This suggests that there were other sources for the paranitrophenol in the urine than exposure to methyl parathion. Indeed, several other common chemicals produce this same metabolite that was used as an indicator of exposure to methyl parathion. These chemicals include acetaminophen, shoe polish, furniture polish, floor polish, leather dressings, paint solvents, gun bluing, metal polishes, scented soaps, spray paints, anything with almond essence, perfumes, and the hydrolysis product of ethyl and methyl parathion in food. Although urine was collected for 48 h prior to workers being actively exposed to methyl parathion, only the amount of paranitrophenol measured in the urine during the 24 h immediately before the exposure activities was used to establish the baseline. The baseline paranitrophenol is the amount excreted in the urine that is unrelated to methyl parathion exposure. The first 24-h urine collections in the hotel rooms (prior to exposure activities) may have contained amounts of paranitrophenol related to previous activities involving exposure to methyl parathion. Although baseline levels of the metabolite, paranitrophenol, were subtracted from the “exposure sample,” the fact that there were detectable levels in the urine during the 24 h preceding the studies’ activities increases the uncertainty in those estimated exposures. EPA considered the risks of exposure from both the microencapsulated formula and the EC formulations. The only formulation of methyl parathion registered in California was the microencapsulated formula. Consequently, only the exposures and risks for workers from the microencapsulated formula are presented for the purposes of comparison in Table 10.2.

10.4.1.3  Bystander Exposure Bystander exposures to methyl parathion are principally through the inhalation route. Air monitoring studies were conducted near application sites for the current major use of methyl parathion on walnut groves (Wofford, 2003). Sampling stations were set up at various distances surrounding walnut groves that were to be treated with methyl parathion. Low-volume air sampling pumps were used to collect 12-h air samples at a flow rate of approximately 1.5 l per minute. Each sampler was fitted with duplicate cartridges containing XAD-2 resin adsorbent. The data served as the primary basis for CDPR’s estimation of bystander exposures to application site air levels of methyl parathion (Formoli, 1999). EPA did not consider bystander exposures to methyl parathion (U.S. EPA, 2006a). The assessment of bystander exposure to a potential toxic air contaminant rests on many assumptions. None of the monitoring data were collected at actual home sites adjacent to treated walnut groves. Consequently, the distance

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Table 10.2  Acute (Single Day) Occupational Exposures to Methyl Parathion Work task

CDPR ADDa (g/kg-day)

CDPR MOEb

EPA ADDc

26.4e

EPA MOEd

Handlers   Mixer/loader (aerial)   Pilot (open cockpit)   Ground mixer/loader

1

–

100

307.0

f

1

–

4

f

1

–

170

f

1

–

5

15.5

e

2

–

75

  Cotton scouts

39.4

1

–

6

  Corn harvesters

11.9

2

–

5

  Walnut rakers

1.3

19

–

14

  Walnut sweepers

0.31

81

–

14

  Walnut shakers

None

100

–

14

  Airblast applicator   Ground boom applicator

265.0 96.0

Re-entry workers

a

Absorbed daily dose (ADD). Margin of exposure (MOE) was defined as acute NOEL/ADD; the acute oral NOEL of 0.025 mg/kg was based on plasma, RBC, and brain ChE inhibition and neuropathology in rats (Koshlukova and Reed, 2006). c Not given by EPA in the IRED (U.S. EPA, 2006a). At the time, chemical-specific data were not available and EPA relied on the 50th percentile of PHED exposure data. For the purposes of comparison, a maximum application rate of 1 lb/acre was used. EPA used a maximum of 350 acres application for aerial, and 200 acres for ground boom. d MOE based on an estimated no effect level of 0.1 mg/kg-day from a subchronic dermal toxicity study in rats. e Calculated from biological monitoring studies data for mixer/loaders and applicators wearing work clothing, coveralls, gloves, shoes, headgear, eyewear, and respirator during application. f Based on pilot, ground mixer/loader, and airblast applicator treating a maximum of 500, 100, and 50 acres in an 8-h workday, respectively (Haskell, 1998) at an application rate of 1 lb of a.i./acre. Corrected for label PPE requirement of a closed cab or coveralls, gloves, shoes, headgear, and eyewear for ground applicators (providing 90% exposure protection). 90% upper confidence limit of the 95th percentile data from PHED were used for acute exposure. b

that such a residence would be from a treated grove is a matter of uncertainty. CDPR assumed that the distance would be minimal, as the air concentrations of methyl parathion and methyl paraoxon that were used in the calculations were indistinguishable from those at the edge of the grove. Likewise, the locations of small children and adults during the periods of time they would have spent outdoors are also matters of uncertainty. CDPR assumed that all individuals exposed to outdoor concentrations of methyl parathion breathed the maximum concentrations of methyl parathion and methyl paraoxon for the entire time. Yet, field-monitoring data indicate there can be a steep drop-off in air concentrations of methyl parathion as the distance from the treated area increases. For example, in one monitoring study there was no discernible air concentration of methyl parathion at 75 yards (69 m) from the grove. It should also be noted that a home has dimensions, and not all portions of a house will be located precisely where the highest levels of methyl parathion and methyl paraoxon were measured. The quantitative effect of building dimensions on the calculated exposures of individuals to indoor levels of methyl parathion and methyl paraoxon cannot be enumerated, but it adds to the uncertainty of the estimates.

There can be a substantial difference between indoor and outdoor air concentrations of a pesticide (Oshima et al., 1981; Segawa et al., 1991). If a structure were to remain closed during the first 24 h after application, it is likely that the residents would have substantially less exposure than if the structure were open. This assertion could be based on using malathion as a surrogate pesticide air contaminant because (1) it has a vapor pressure (4  105 mm Hg at 30°C) similar to methyl parathion (1.7  105 mm Hg at 25°C); (2) the malathion formulation has a noxious odor (due to inerts, and thiols, sulfides and disulfides produced in the manufacturing process) like the methyl parathion formulation so it would be reasonable to assume that windows and doors would be closed as much as possible; (3) the monitored surrogate homes and structures were located in the areas sprayed directly with malathion; and (4) extensive measurements of outdoor and indoor air concentrations of malathion vapor were conducted. The fact that the homes and other structures monitored for malathion were located in an urban area, rather than on a farm (Oshima et al., 1981; Segawa et al., 1991), would contribute a degree of uncertainty. However, the fact that more than 80 structures were monitored would tend to

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take potential architectural differences between urban and farm structures into consideration. Nevertheless, nothing on the label, or in California regulations or permit conditions, requires bystanders to close up their homes during pesticide applications. Thus, both indoor and outdoor air concentrations of methyl parathion and methyl paraoxon could be the same. Given this assumption, the estimated absorbed daily doses of methyl parathion and methyl paraoxon for bystanders living adjacent to application sites are based on the highest measured air concentrations of those two chemicals (Table 10.3). The reported methyl paraoxon values are probably overestimates because artificial oxidation of methyl parathion to methyl paraoxon occurs in the sampling device in the technique that was used (Segawa et al., 1991). However, the amount of oxidation that occurs during the sampling process cannot be determined with accuracy, so this adds to the uncertainty of the estimated ambient air concentrations of methyl paraoxon. Human exposure was calculated as an absorbed dosage. The absorbed dosage per unit of body weight varies between infants, children, adult females, and adult males because the ratio of inhalation rate to body weight varies from one subgroup to another. Therefore, the estimate of human exposure is separated into these four subgroups. Infants of age 6 months were chosen because they usually are the highest exposure subgroup due to the highest inhalation rate to body weight ratio (Andrews and Patterson, 2000).

10.4.2  Methyl Bromide Methyl bromide (MB, CAS #74–83–9) is a colorless gas, usually odorless, with a sweetish, chloroform-like odor at high concentrations (odor threshold at 80 mg/m3 or 20.6 ppm) and a burning taste. Methyl bromide has been widely used as a fumigant to control pests in soil prior to planting, to fumigate fresh and dry agricultural products, and to fumigate residences and commercial structures.

Table 10.3  Maximum, Acute Bystander Exposures and Risks from Airborne Methyl Parathion and Methyl Paraoxon Measured at 10 Yards (7.5 m) from the Application Site in CDPR Studies Individual

Methyl parathion ADDa (g/ kg-day)

Methyl paraoxon ADDb (g/ kg-day)

CDPR MOEc

Infant

2.59

0.172

6

Child, 3–5 years

2.02

0.134

7

Adult female

0.76

0.051

20

Adult male

0.93

0.061

16

a

Absorbed daily dose (ADD) based on a 24-h time-weighted average methyl parathion air concentration of 4.38 g/m3, CDPR default values for breathing rates and body weights (Andrews and Patterson, 2000), assuming 100% inhalation retention and absorption. b ADD based on a 24-h time-weighted average methyl paraoxon air concentration of 0.29 g/m3. c Margin of exposure (MOE) was defined as acute NOEL/ADD; the acute oral NOEL of 0.025 mg/kg was based on plasma, RBC, and brain ChE inhibition and neuropathology in rats (Koshlukova and Reed, 2003). Assumes paraoxon is 10 times more toxic than parathion.

methyl bromide (U.S. EPA, 2006b). EPA assumed that the developmental anomalies could occur as a result of a single exposure. There was an acute, whole-body inhalation, neurotoxicity study conducted in rats that had a single-dose LOAEL of 350 ppm for clinical signs and FOB effects (NOAEL  100 ppm; 388 g/m3) (Driscoll and Hurley, 1993). However, EPA decided to use the lower NOAEL from the developmental toxicity study because it was “more health protective” (U.S. EPA, 2006b). Rather than calculate acute absorbed doses of MB for laboratory animals and humans in the RED, EPA used human equivalent concentrations (HECs) for nonoccupational and occupational risk assessment. These HECs were calculated by using the formula

10.4.2.1  Toxicology The 4-h LC50 for methyl bromide in rats was 780 ppm (Lim, 2002; U.S. EPA, 2006b). Methyl bromide causes severe skin and eye irritation in humans (Alexeef and Kilgore, 1983; Hezemans-Boer et al., 1988). EPA selected a no-observed-adverse-effect level (NOAEL) of 40 ppm (155.2 g/m3) from a developmental toxicity study using whole-body inhalation exposure of rabbits [lowest observed adverse effect level (LOAEL)  80 ppm for developmental anomalies (agenesis of the gall bladder and increased incidence of fused sternebrae); Breslin et al., 1990] as the basis to gauge the acute risks of inhalation exposure to air concentrations of





where POD  point of departure (NOAEL) in the critical toxicology study, Dl.a. expos.  duration of laboratory animal exposure (h/day), Dhum. expos.  duration of human exposure (h/day), W1  number of days laboratory animals exposed, W2  number of days humans exposed, and RGDR  regional gas dose ratio. The regional gas dose ratio, related to the ratio of the minute volume to the surface area of the affected portions of the respiratory tract in rabbits compared to humans, was assumed to be 1 for methyl bromide because it was “health protective” (U.S. EPA, 2006b). The actual RGDR is

Chapter | 10  Risk Assessment for Acute Exposure to Pesticides

probably much less. The absorbed dose of a rabbit breathing methyl bromide at 40 ppm (155.2 g/m3) for 24 h is 81.6 g/kg-day. The absorbed dose of an adult human breathing the same concentration of methyl bromide for a day is 44.3 g/kg-day, or 54% of the rabbit dose. No studies were located that compared whole-body to nose-only inhalation studies in rabbits, as has been done in rats (Blair et al., 1974; Hext, 1991; Iwasaki et al., 1987; Jaskot and Costa, 1994; Landry et al., 1986; Langard and Nordhagen, 1980; Tyl et al., 1995; Wolff et al., 1982). Because methyl bromide is an irritant (Lim, 2002; U.S. EPA, 2006b), it is likely that grooming behavior would have added a substantial nondietary oral component to the amount absorbed through the inhalation route in the rabbit developmental toxicity study. That would further skew the rabbit to human ratio of absorbed dose. EPA appeared to try to compensate for this by using an interspecies extrapolation factor of 3, instead of 10, explaining that “the RfC methodology takes into consideration the pharmacokinetic differences but not the pharmacodynamic differences” (U.S. EPA, 2006b). CDPR also used the rabbit developmental toxicity study NOEL of 40 ppm as the critical endpoint to gauge the risks of acute exposures to MB (Lim, 2002). Similar to EPA, CDPR used human equivalent air concentrations of MB for nonoccupational and occupational risk assessment. However, CDPR took the differences in breathing rates into consideration using this equation to modify the NOEL:





The respiratory rates were 0.54 m3/kg-day for rabbits and 0.26 m3/day for adult humans (Zielhuis and van der Kreek, 1979). Consequently, the calculated human equivalent NOEL for a 24-h exposure was 21 ppm. This number was used to gauge the risks of acute adult exposures to methyl bromide for resident bystanders as well as handlers. (a)  Pharmacokinetics Radiolabeled methyl bromide was rapidly biotransformed and readily excreted in rats following inhalation (Bond et al., 1985). In all tissues examined, over 90% of radioactivity was in the form of metabolites. The elimination half-life of radioactivity from tissues was 1.5–8 h. Almost 50% of the absorbed dose was excreted via the lungs as CO2. The pulmonary excretion was biphasic, with halflives of 3.9  and 11.4 h. The half-lives of radioactivity in the urine and feces were 9.6  and 16.1 h, respectively. In a different rat inhalation study (Bond et al., 1985), the percentages of the absorbed dose excreted in the urine and feces were 23 and 2%, respectively. In other rat inhalation studies, Medinsky et al. (1984) and Jaskot

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et al. (1988) observed similar results with regard to exhalation and excretion of the absorbed dose, as well as excretion half-lives. In a human study, the amount of 14C-MB metabolized and exhaled as 14CO2 ranged from 0.2 to 1.0% of the dose for mouth breathing, and 0.2 to 0.4% of the dose for nose breathing. Measurements were conducted at the end of 2 h of exposure and 2 h of clearance (Raabe, 1988). The net body retention for both exposure routes was 51.1% with a clearance half-life of 72 h, based on the amounts of radiolabel in exhaled air and the urine 0.5 h after inhalation exposure. The inhalation route is not the only possible route of exposure to pesticide vapors. Pesticide vapors come in contact with the skin. It was reported that one fatal and two nonfatal cases of poisoning occurred after the fumigation of a flour mill many years ago (Jordi, 1953). Results of the investigation revealed that the workers had worn oxygen-supplying apparatus, with adequate oxygen during the 1.5-h fumigation period. All workers experienced illness symptoms at least 1 h after the fumigation. However, no dermal absorption studies were submitted to CDPR. Consequently, the amount of methyl bromide absorbed through the dermal route cannot be quantified accurately. Although the amount absorbed through the skin is likely to be substantially less than the amount retained/ absorbed through the lungs, to be health protective there should be some indication of the significance of the dermal contribution. Examination of the literature suggested a possible approach to obtaining a theoretical estimate of dermal absorption of methyl bromide. In general, the permeability of a chemical through skin is related to the chemical’s partitioning into air, blood, and lipids (McDougal et al., 1990; U.S. EPA, 1992a). Mattie et al. (1994) determined skin–air partition coefficients for several volatile organic chemicals in an in vitro study using clipped, whole-thickness rat skin and compared these partition coefficients with octanol– water partition coefficients reported by Leo et al. (1971) and rat skin permeability reported by McDougal et al. (1985, 1986, 1990). Mattie et al. (1994) found that skin–air partition coefficients correlated well with skin permeability (r2  0.93) but that octanol–water partition coefficients did not (r2  0.09). In its guidance for estimating dermal exposure, U.S. EPA (1992a) suggests that the fat–air partition coefficient for an airborne chemical may be used to estimate skin permeability. The formula, suggested by U.S. EPA, to make that estimate, is as follows:



where Kp(est)  the estimated skin permeability coefficient, and Kf/a  the fat/air partition coefficient. Methyl iodide, a chemically similar fumigant, may be used as a surrogate for methyl bromide, In the case of methyl iodide, the measured Kf/a in rats is 88.8  2.3

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(Gannon, 2004). Thus, substituting 88.8 for Kf/a in the previous formula yields an estimated Kp of 0.005 cm/h. Dermal absorption of methyl iodide may then be estimated using dermal permeability coefficients, based on Fick’s first law (McDougal et al., 1990):

where Kp  measured or calculated skin permeability coefficient (cm/h), conc.exposure  concentration of the chemical in air (g/m3), areaskin  area of skin exposed (cm2), and timeexposure  duration of exposure period (h). The unmodified air concentration of methyl bromide equivalent to the NOEL for both CDPR and EPA is 155g/m3 (Lim, 2002; U.S. EPA, 2006b). A generic adult is assumed to have a total body surface area of 18,150 cm2 (U.S. EPA, 1997). Thus, the amount of methyl bromide absorbed dermally by applicators in an 8-h period would be

The dose of MB absorbed through the inhalation route by an applicator experiencing the NOEL air concentration for 8 h was estimated to be 44.3 g/kg-day. If we assume the generic adult weighs 70 kg (U.S. EPA, 1997), the amount of methyl bromide absorbed through the dermal route would be 0.0016 g/kg-day. Consequently, the amount theoretically absorbed through the dermal route (0.0016 g/kg) constitutes approximately 0.004% of the amount absorbed through the inhalation route. This amount of exposure is usually considered insignificant (Donahue, 1996).

10.4.2.2  Exposure Estimates Methyl bromide exposure estimates for individuals are derived from the scenarios that involve (1) fumigation of pre-plant soil, (2) fumigation of agricultural commodities or structures, as well as (3) fumigation for residents or persons who live or work at the edge of the buffer zone distances from commodity or field fumigations. Many studies were available to provide data for occupational and bystander exposures associated with pre-plant field fumigations (Thongsinthusak and Haskell, 2002; U.S. EPA, 2006b). The EPA approach to estimating occupational exposures was somewhat different than CDPR’s: 1. In the submitted studies, each worker wore air samplers. If two were worn, CDPR considered the average of the two samplers as a single replicate. EPA used each sampler as a replicate. 2. The application rates used in the studies were different than the maximum application rate on the labels. CDPR

adjusted the exposures to reflect the maximum application rate. EPA did not adjust for the maximum application rate. 3. CDPR used the reported field spike data to make its own adjustment for recovery and analytical technique. EPA used the registrant’s calculated field spike adjustments. 4. CDPR calculated an upper bound for an acute 8-h exposure for workers. In some instances the upper-bound values exceeded the highest measured value. EPA used the maximum measured air concentration of methyl bromide to represent the acute 8-h exposure for workers. A comparison of the 8-h acute air concentrations, estimated by CDPR and EPA is presented in Table 10.4.

10.4.2.3  Bystander Exposure In order to estimate bystander exposures adjacent to fields treated with methyl bromide, it is necessary to obtain an estimate of the methyl bromide air concentrations. CDPR did not use the direct sampling method for estimating application site air concentrations because there are several uncertainties associated with the use of the direct sampling method that limit its utility. Instead, air concentrations of methyl bromide were first measured by fixed samplers that were positioned at various locations around the treated area (both downwind and upwind, as well as at points in between). As a rule, air concentrations of fumigants measured downwind tend to be relatively high, as a

Table 10.4  Comparison of Acute Occupational Exposures to Methyl bromide in Pre-Plant Field Fumigation as Estimated by CDPR (Thongsinthusak and Haskell, 2002) and EPA (U.S. EPA, 2006b) Work task

CDPR estimated exposure conc. w/label-approved PPE (ppm)a

EPA estimated exposure conc. w/label-approved PPE (ppm)a

Shank injection-tarped soil fumigation (broadcast and bedded)  Tractor driver (shallow shank) 0.111

0.10

  Co-pilot

0.224

0.08

 Tractor driver (deep shank)

0.38

0.11

  Shovelman

0.02

0.007

  Tarp cutter

0.001

0.0007

Tarped-bed fumigation drip irrigation   Applicator a

0.025

0.024

Expressed as a 24-h time weighted average but assumes 8-h workday.

Chapter | 10  Risk Assessment for Acute Exposure to Pesticides

fumigant plume will be pushed by the wind in that direction. Concentrations of fumigant upwind tend to be low, or close to zero, as a plume will be pushed by the wind in the opposite direction. Thus, there can be a very large difference between upwind and downwind air concentrations of a fumigant. In areas where there is a predominant wind direction, averaging of the air concentrations from these various samplers is probably not appropriate as persons around treated areas will generally be in one location relative to the wind. Consequently, they will not be exposed to an average of these concentrations. Second, samplers were positioned at specific distances from the treated area, and the measured concentrations represent air concentrations only at those distances. As air concentrations vary greatly by distance, the air concentrations estimated from direct measures represent a very narrow range of the possible levels to which people can be exposed. Finally, the measured air concentrations represent only those for the conditions under which the studies were carried out. Air concentrations around treated fields, buildings, or other areas are influenced by a number of factors, including how a chemical is applied, application rates, techniques designed to control emissions (e.g., tarps), and weather conditions. Varying weather conditions, for example, can significantly change the air concentrations at specific sites around a treated area. As there is a large range of potential weather conditions that can exist, it is not possible for these studies to represent the entire range of potential exposures that can result from different weather conditions. Screening level modeling with the Industrial Source Complex Short Term (ISCST version 3) model produces reasonable worst case estimates of air concentrations and resulting risks for a number of reasons. First, only downwind center-line air concentrations expected under reasonable worst case meteorological conditions for a particular averaging time scenario are considered. Thus, the screening level air concentrations estimated by the ISCST3 model would be found in the upper percentiles of air concentration distributions obtained from using historical weather data. However, the model does allow for estimation of air concentrations that reflect different conditions based on changing factors, such as application rate, field size, downwind distances, and weather conditions. These factors cannot be taken into account by using monitoring data alone. Consequently, the ISCST3 screening level results should be considered to represent potential exposures to the most highly exposed, upper percentile of the population. However, those results are not representative of exposures to most of the population situated around a treated field. When all other factors are held constant, the ISCST3 model uses an equation that makes the flux and the air concentrations directly proportional. A number of factors may affect the flux of methyl bromide from the fields

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where it has been applied. These factors contribute to the uncertainty in the estimates of the air concentrations near application sites. Soil: Field study results for other fumigants support the use of water applications to suppress flux by increasing soil moisture. Flux may also be a function of soil textures and temperatures. However, DPR does not have studies that adequately quantify the magnitude of the effect of those factors. Farming technique: Generally, tarped soil shows lower flux than untarped soil. However, the magnitude of this effect depends on both the fumigant and the type of tarp used. Field study results indicate that tarped raised-bed applications show higher flux than tarped broadcast applications. Sometimes additives are used to fertilize the soil during drip irrigation applications. These additives may interact with the fumigant to change the fumigant flux. Another area of uncertainty concerns the relationship between flux, concentration, and meteorological conditions. Flux is usually lower at night. However, several field studies demonstrate that for some fumigants and/or application methods the highest flux occurs at night. Regardless of the magnitude of flux, air concentrations tend to be highest at night due to the very stable atmospheric conditions that are characteristic of nighttime hours. Thus, nighttime flux may result in very high air concentrations even though that night flux appears to be relatively small in magnitude compared to daytime flux values. Atmospheric stability in this case refers to the degree of vertical atmospheric mixing. Atmospheric conditions during the day tend to be much less stable relative to night conditions. Vertical mixing during the day is increased due to heating of the earth’s surface. Any pollutants in the air are diluted as they are mixed upward into clean air. This leads to generally lower air concentrations of methyl bromide during the day. Air dispersion modeling defines night as the period from 1 h before sunset to 1 h after sunrise. Atmospheric conditions during night tend to be stable to very stable (cold, dense air near the soil: warmer, lighter air at greater heights, little or no vertical mixing). Calm winds are associated with stable atmospheric conditions at night. Inversion conditions may also (but not always) be present. Under calm wind conditions, there is little or no horizontal (crosswind) spreading of a pollutant plume. Pollutant plumes tend to stay intact and concentrated for great distances beyond the source edge when there is little vertical or horizontal dilution of the pollutant plume under these calm wind and stable atmospheric conditions. Thus, even if flux is lowest at night, nighttime stable conditions can lead to very high air concentrations. The location of the highest off-site air concentrations is uncertain because the crosswind direction movement of the pollutant plume under calm winds is erratic and unpredictable. These factors cause air concentrations associated with fumigants to be highest at night. Several large

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residential fumigant exposure incidents have occurred under nighttime conditions, particularly at or shortly after sunset. Finally, air concentration is proportional to flux in the Gaussian plume model. DPR also assumes that flux is proportional to application rate but that flux does not vary with application size (Segawa, 1997). These assumptions together permit the use of the ISCST3 model to estimate off-site air concentrations for application sizes other than those directly monitored. (a)  Buffer Zones A method used to mitigate human health hazards associated with the use of soil fumigants is to require buffer zones around the fumigated area. The idea is to prevent people from inhabiting the buffer zone areas where the concentrations may be higher than a reference concentration. In the case of methyl bromide, the reference concentration is 815 g/m3 for 24 h. CDPR promulgated flux- and area-dependent buffer zones to mitigate the acute concentrations (Johnson, 2001). This mitigation depends on several factors in order to be effective. Concentrations must decline as a function of distance from the field. This is generally true for single, ground-level, area sources. The mitigation also requires that concentration variations are not so great as to exceed the reference concentration in the buffer zone. The previous discussion deals with the factors that affect the variability in concentrations that occurs both over space and over time. EPA used both the ISCST3 model and the probabilistic exposure and risk model for fumigants (PERFUM) to evaluate distributional bystander exposure from data derived from fumigation studies conducted in California and elsewhere (Reiss and Griffin, 2006; U.S. EPA, 2008). As ISCST3 is an integral part of the PERFUM model, many of the inputs used for PERFUM are similar to those used for the ISCST3 analysis (e.g., field sizes and backcalculated flux rates). The key difference is that PERFUM incorporates 5 years of meteorological data to generate a distribution of daily average concentrations that represent the possible range of downwind air concentrations based on changing wind vectors from the measured data in a series of receptor locations. The EPA Science Advisory Panel (SAP) concluded in their review that, in concept, the PERFUM model was reasonable. However, the SAP did not perform an in-depth assessment of the reliability of the PERFUM front and back end processing code as it was not their charge. CDPR has made a practice of thoroughly evaluating air dispersion models before utilizing them in risk assessment. Although the ISCST3 model had been thoroughly evaluated at CDPR, the new PERFUM components had not. Therefore, only screening level air concentration estimates were used for the CDPR MB exposure assessment.

Hayes’ Handbook of Pesticide Toxicology

In a later iteration of a draft risk assessment for methyl bromide (U.S. EPA, 2008), EPA estimated “whole field” buffer zone distances near 40-acre fields using the PERFUM model and “target concentrations” derived from various acute toxicological endpoints. The EPA buffer zone distances were expressed as the distance from the edge of a treated field to a point chosen at random where there was a 99% probability that the time weighted average (TWA) air concentration of MB would be less than or equal to a target concentration. A target concentration was defined as that air concentration of methyl bromide which, when divided into a toxicological NOAEL from a laboratory animal study, yielded a number equal to or greater than the appropriate uncertainty factors. This whole-field, probabilistic approach differs from CDPR’s maximum direction approach (Barry and Johnson, 2008). The two approaches were compared using air concentration data from 20-acre field fumigations with methyl bromide (24-h TWA), metam sodium (8-h TWA), and chloropicrin (4-h TWA). With each set of data, the PERFUM model was used to establish the whole-field buffer zones where any random point on the periphery had a 99% probability that the fumigant air concentration would be equal to or less than a target concentration. The PERFUM model was also used for the maximum direction approach for each of these fumigants. This latter analysis indicated that at the whole-field buffer zone distance some points on the periphery had greater than a 1% probability of having a fumigant air concentration greater than the target concentration. The portion of the 99% whole-field buffer zone perimeter along which the exceedance (failure rate) occurred could be as long as a football field. In the case of methyl bromide (24-h TWA), the failure rate ranged from 12 to 14%. For metam sodium (8-h TWA), the failure rate was 7.5–22%, and in the case of chloropicrin (4-h TWA) the failure rate ranged from 10 to 29%. Finally, there are default factors used by CDPR and EPA that affect the size of the buffer zones. The reference level for methyl bromide used by EPA is 1294 g/m3, while CDPR uses 815 g/m3. This tends to make CDPR’s buffer zones longer. However, EPA uses the 99.9th percentile of exposure, while CDPR uses the 95th percentile. This tends to make EPA’s buffer zones longer. In the case of tarped, shallow broadcast applications, the factors interact to make EPA’s buffer zones longer at the lower emission ratio. In the case of tarped, bedded applications, the factors interact to make CDPR’s buffer zones longer at the higher emission ratio. Some of the CDPR estimates of occupational exposures to methyl bromide resulting from commodity and commercial structural fumigations are shown in Table 10.5. Buffer zones around commodity and industrial structure fumigation sites were set by CDPR using a 15-min TWA

Chapter | 10  Risk Assessment for Acute Exposure to Pesticides

Table 10.5  Air Concentrations of MB Associated with a Specific Job Task, as Measured in a Worker’s Breathing Zone Type of fumigation Work task

CDPR estimatea 8-h TWA (ppm)

Flour mill—gas tanks Applicator inside building

12.9

Aerator

21.6

Tape removers

0.4b

Flour mill—gas tanks Applicator outside building

Processing and handling silo— enclosed conveyer and storage bins

Grain silo, elevator, or bin

Shipping containers—trailers or rail cars

Tarpaulin—wooden furniture and pallets of flour

0.7

Aerator

20.1

Applicator

17.3

Aerator

0.1b

Applicator

0.7

Aerator

0.2b

Grain loader

0.2b

Applicator

0.02b

Aerator

6.8b

Applicator

0.1b

Tarp remover

0.2b

Aerator

0.1b

Applicator Flat storage building—loose corn, soybeans

0.5

Helper

0.1b

Aerator

0.1b

a 95th percentile of exposure based on data from table 10 in the CDPR exposure assessment document (Thongsinthusak and Haskell, 2002). b Not enough data to calculate 95th percentile of exposure.

of methyl bromide air concentrations and adjusting the toxicological endpoint (developmental toxicity from 6-h daily exposures of rabbits; Breslin et al., 1990) accord­ in­gly (Barry, 2005; Lim, 2002). This approach causes a great deal of uncertainty, as it is totally unclear that a single, 15-min exposure to methyl bromide would cause

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developmental anomalies in a human. The derivation of EPA’s buffer zones around commodity and industrial structure fumigation sites was unclear (Barry, 2005, 2006).

Conclusion The toxicological database for pesticides as a chemical use class is the most complete for any category of chemicals, with the possible exception of pharmaceuticals. The FIFRA- and EU-required toxicological studies are designed to produce data on the types of toxic effects a chemical might possibly cause, as well as the dose that is required to elicit that effect. Nonetheless, these studies often do not contain the type of information needed for accurate estimates of the acute risks associated with the use of a pesticide. This situation can be complicated by political intrusions into scientific databases, such as the executive order not to consider human toxicity data (U.S. EPA, 2007a). only one of the currently required FIFRA guideline studies examines the amount of chemical needed to produce nonlethal systemic effects from a single dose. The limited availability of applicable data sometimes forces the critical (regulatory) acute NOELs to be derived from multiple-dose laboratory animal studies, such as developmental toxicity studies. In those instances where the assessor uses NOELs for systemic effects caused by multiple doses, as in the examples of methyl parathion and methyl bromide, there can be an overstatement of the risks from a single dose. Most human exposure to nonfumigant pesticides occurs through the skin (Ross et al., 1992; Wolfe, 1976). Yet, almost all of the required FIFRA guideline studies utilize the oral exposure route for dosing for nonfumigants. Clearly, the respective types of data do not fit together well. Dermal pharmacokinetic studies, which provide data on plasma levels and the half-lives of pesticides and their metabolites, could be used to better understand the differences in toxicokinetics for chemicals entering the body through the skin rather than via the digestive system. Currently, however, such studies are not required. One of the major issues in selecting a critical endpoint for assessing the risks of acute exposure to pesticides is recognizing the difference between indications of exposure and an adverse effect. This theme carries through some of the examples presented in this chapter. Perhaps one of the most controversial, though extremely well-researched pesticides, is chlorpyrifos. The toxicological endpoints selected by EPA and CDPR were very different, and so, concomitantly, are the estimated risks (Cochran, 2002). This controversy extends even beyond the regulatory agencies. A nongovernmental group considered reduced RBC AChE activity to be the critical effect, and thus would use the multiple-dose, oral human NOEL of 0.1 mg/kg-day from

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the Coulston study (Coulston et al., 1972) to gauge the risk of acute toxic effects from chlorpyrifos (Zhao et al., 2006). EPA, in the Guidelines for Neurotoxicity Risk Assessment, lists alterations in enzyme activity causing the degradation of neurotransmitters as a possible adverse effect because it can lead to unwanted changes in the function of the nervous system (U.S. EPA, 1998e). Reduced levels of RBC AChE activity do not affect the level of acetylcholine at either synaptic or neuromuscular junctions. Reduced levels of ChE activity in the blood are indications of exposure, not adverse effects. It should be noted that the guidelines do not specify the level of inhibition of even brain AChE activity which constitutes an adverse effect. Studies have shown that statistically significant reductions in brain AChE activity caused by chlorpyrifos may not lead to cholinergic signs in laboratory animals (Bushnell et al., 1993, 1994; Chanda and Pope, 1996; Stanton et al., 1994). However, organophosphorus insecticide poisoning in humans may lead not only to cholinergic signs but also to symptoms, such as headaches and nausea (Ellenhorn et al., 1997), which cannot be ascertained in laboratory animals. Consequently, statistically significant reduced levels of brain AChE activity in laboratory animals are generally used as a surrogate for this manifestation (clinical symptoms) of an impaired nervous system function (JMPR, 1998; U.S. EPA, 1998e). The basic difficulty in deriving a critical NOEL (adverse effect) to use for gauging risks of acute exposures to pesticides is that regulators do not want to underestimate those risks. Consequently, toxicologists search for the most sensitive toxicological endpoint documented in FIFRA guideline studies or in the published literature. This desirable focus on being health protective during endpoint selection causes difficulties in discriminating between toxicological signals and noise. In addition, to be health protective, regulatory agencies assume that the least sensitive human is 10 times more sensitive to the toxic effects of a pesticide than the most sensitive laboratory animal (Davidson et al., 1986; Dourson and Stara, 1983, 1985; U.S. EPA, 1986). Further, both EPA and CDPR assume that the range of sensitivity within humans is also 10-fold (Burin and Saunders, 1999). On the exposure side of risk assessment, although there are guidelines that exposure studies must meet, there are no required studies comparable to the required toxicity studies. Many of the exposure studies that have been done are outdated with respect to equipment, methods of use, or they have been incorporated into databases such as PHED. It is not feasible to produce the myriad of exposure monitoring studies that would be required to fully describe the exposures to a pesticide that arise from differences in formulations, the various methods in which those formulations can be used, and the resultant incidental exposures of the public. Consequently, both the amount and the quality

of data available on which to base exposure estimates are highly variable. The uncertainties in the data result in a commensurate number of conservative (health protective) assumptions. Registrants, agribusiness, and the pesticide control industry often complain that these conservatisms are simply piled one upon another until the estimated exposure for a scenario bears no resemblance to reality. Qualitatively, these complaints are probably correct. However, government regulators are charged with protecting public health, and with that responsibility is an obligation to avoid unnecessary risks. As regulatory estimates of exposure must be based on logical, defensible calculations from the best data available, government scientists, faced with uncertainties from inadequate data, prefer to err on the side of safety. With better data there is less uncertainty, and the accuracy of the exposure estimates is increased.

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U.S. Environmental Protection Agency (EPA) (1991). “Guidelines for Developmental Toxicity Risk Assessment,” CFR 56 (No. 234): 6379863826. U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (1992a). “Dermal Exposure Assessment: Principles and Applications (Interim Report),” EPA/600/8–91/01 1B. Office of Research and Development, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (1992b). “Guidelines for Exposure Assessment: Notice,” EPA/600/Z-92/001. U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (1996). “Exposure Factors Handbook,” EPA/600/8–89/043. Office of Health and Environmental Assessment, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (1997). “Series 875— Occupational and Residential Exposure Test Guidelines. Group B Postapplication Exposure Monitoring Test Guidelines.” Office of Prevention, Pesticides, and Toxic Substances, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (1998a). “Health Effects Test Guidelines OPPTS 870.1000 Acute Toxicity TestingBackground,” EPA 712-C-98–189. Office of Prevention, Pesticides, and Toxic Substances, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (1998b). “Health Effects Test Guidelines OPPTS 870.6200 Neurotoxicity Screening Battery,” EPA 7 12-C-98–238. Office of Prevention, Pesticides, and Toxic Substances, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (1998c). “Health Effects Test Guidelines OPPTS 870.7485 Metabolism and Pharmacokinetics,” EPA 712. C-98–244. Office of Prevention, Pesticides, and Toxic Substances, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (1998d). “Framework for Assessing Non-occupational, Non-dietary (Residential) Exposure to Pesticides.” U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (1998e). “Guidelines for Neurotoxicity Risk Assessment; Notice.” Fed. Reg. 63(93), 26925– 26954. Thursday, May 14, 1998. U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (1999). “Use of Values from the PHED Surrogate Table and Chemical-specific Data,” Policy Number 007. Science Advisory Council for Exposure, U. S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (2001a). “Recommended Revisions to the Standard Operating Procedures (SOPs) for Residential Exposure Assessments,” Policy 12. U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (2001b). “Standard Values for Daily Acres Treated in Agriculture,” Policy Number 009.1. Science Advisory Council for Exposure, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (2004). “MGK-264: Occupational and Residential Exposure Assessment for the Reregistration Eligibility Decision Document.” Health Effects Division, Office of Pesticide Programs, U. S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (2006a). “Interim Rere­ gistration Eligibility Decision for Methyl Parathion. Memorandum from Debra Edwards to Jim Jones,” July 31, 2006. Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC.

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U.S. Environmental Protection Agency (EPA) (2006b). “Human Health Risk Assessment: Methyl Bromide.” Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (2007a). “Expanded Protections for Subjects in Human Studies Research.” Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. http://www.epa.gov/oppfead1/guidance/human-test.htm. U.S. Environmental Protection Agency (EPA) (2007b). “Review of Worker Exposure Assessment Methods.” Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. http://www.epa.gov/osa/hsrb/files/meeting-materials/apr-18-20-2007public-meeting/ReviewOfWorkerExposureAssessmentMethods.pdf. U.S. Environmental Protection Agency (EPA) (2008). “Reregistration Eligibility Decision (RED) for Methyl Bromide.” Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. van Hemmen, J. J. (1992). Estimating worker exposure for pesticide registration. Rev. Environ. Contam. Toxicol. 128, 43–54. Wester, R. C., and Maibach, H. I. (1976). Relationship of topical dose and percutaneous absorption in rhesus monkey and man. J. Invest. Dermatol. 67, 518–520. Wester, R. C., and Maibach, H. I. (1985). In vivo percutaneous absorption and decontamination of pesticides in humans. J. Toxicol. Environ. Health 16, 25–37. Wester, R. C., and Maibach, H. I. (1993). Animal models for percutaneous absorption. In “Health Risk Assessment: Dermal and Inhalation Exposure and Absorption of Toxicants” (R. G. M. Wang, J. B. Knaak, and H. I. Maibach, eds.), pp. 89–103. CRC Press, Boca Raton, FL. Willard, T. R. (2000a). “Biomonitoring Assessment of Worker Exposure to Methyl Parathion during Cotton Scouting Following Applica­tions of Penncap-M Microencapsulated Insecticide.” Unpubli­shed report. Cerexagri Study No. KP-2000–02. CDPR, Vol. No. 121–155, #185862. California Environmental Protection Agency, Sacramento, CA. Willard, T. R. (2000b). “Biomonitoring Assessment of Worker Exposure to Methyl Parathion during Sweet Corn Hand-Harvesting Following Applications of Penncap-M Microencapsulated Insecticide.” Unpublished report. Cerexagri Study No. KP-99–17 CDPR, Vol. No. 121–154, #185861. California Environmental Protection Agency, Sacramento, CA. Willard, T. R. (2001). “Occupational Exposure of Monitoring of Mixing/Loading Activities for Aerial Application of Penncap-M Microencapsulated Insecticide.” Unpublished report. Cerexagri Study No. KP-2001–03, CDPR Vol. No. 121–152, #185859. California Environmental Protection Agency, Sacramento, CA. Wofford, P. (2003). “Preliminary Methyl Parathion Air Concentrations Measured Adjacent to Two Walnut Orchard Applications.” Unpublished report. Environmental Monitoring and Pest Management, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, CA. Wolfe, H. R. (1976). Field exposure to airborne pesticides. In “Air Pollution from Pesticides and Agricultural Processes” (R. E. Lee, ed.), pp. 137–161. CRC Press, Boca Raton, FL. Wolff, R. K., Griffis, L. C., Hobbs, C. H., and McClellan, R. O. (1982). Deposition and retention of 0.1 g aggregate aerosols in rats following whole body exposures. Fundam. Appl. Toxicol. 2, 195–200. Woollen, B. (1993). Biological monitoring for pesticide absorption. Ann. Occup. Hyg. 37, 525–540. World Health Organization (WHO) (1978). “Environmental Health Criteria 6. Principles and Methods for Evaluating the Toxicity of Chemicals. Part I.” WHO, Geneva.

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World Health Organization (WHO) (1984). “Pesticide Residues in Food. FAO Plant Production and Protection Paper No. 67,” pp. 685–688. Food and Agriculture Organization and the World Health Organization, Geneva. Zendzian, R. P. (1994). “Dermal Absorption of Pesticides. Pesticide Assessment Guidelines. Subdivision F, Hazard Evaluation: Human and Domestic Animals,” Series 85–3. Health Effects Division, Office

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of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. Zhao, Q., Dourson, M., and Gadagbui, B. (2006). A review of the reference dose for chlorpyrifos. Regul. Toxicol. Pharmacol. 44, 111–124. Zielhuis, R. L., and van der Kreek, F. W. (1979). The use of a safety factor in setting health based permissible levels for occupational exposure. Int. Arch. Occup. Environ. Health 42, 191–201.

Chapter 11

Genotoxicity of Pesticides David A. Eastmond and Sharada Balakrishnan University of California, Riverside, California

11.1  Introduction Pesticides are biologically active compounds selected and used for their biocidal properties. In many cases, these agents are highly specific in their toxic effects, acting on a unique molecular target or affecting a narrow range of organisms. However, in other cases, these agents can affect a much broader range of targets and organisms, including humans. As a result, there exist ongoing concerns about the health effects of pesticide exposure in humans. These concerns have been heightened by pesticide-related poisoning episodes that have occurred during the past 50 years, such as those involving hexachlorobenzene (Schmid, 1960), methylmercury (Bakir et al., 1973), malathion (Baker et al., 1978), dibromochloropropane (Slutsky et al., 1999; Whorton et al., 1979), aldicarb (Green et al., 1987), methylparathion (Rehner et al., 2000), and methamidophos (Sumi et al., 2008). In addition to acute effects, substantial concerns exist about chronic effects such as cancer and heritable diseases that might stem from pesticide exposure. An association between pesticide exposure and cancer has been suspected for more than 50 years following reports of the occurrence of elevated levels of skin and lung cancer in European farmers using arsenical insecticides in grape production (Jungmann, 1966; Roth, 1958; Thiers et al., 1967). In a few cases, the association between pesticide exposure and cancer has been confirmed [Blair and Zahm, 1995; Institute of Medicine (IOM), 1999; International Agency for Research on Cancer (IARC), 1987a, 1994; Zahm et al., 1997]. However, in many cases these concerns remain unsubstantiated either due to an underlying lack of an association or because of the difficulties in conducting epidemiological studies in these exposed populations. Even where associations have been seen or suspected, identifying the specific agent responsible has been difficult for a variety of reasons, including poorly defined and variable exposure levels, concurrent exposure to multiple pesticides as well as other potentially carcinogenic agents, long latency periods, small study populations, Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

and other potential confounding factors. In addition, as additional limitations, human epidemiological studies are costly and can only take place following exposure— an approach that is not considered protective of public health. Because of these difficulties, regulatory agencies and other organizations have turned to chronic animal bioassays, short-term tests, and other approaches such as human biomonitoring to supplement conventional human epidemiological approaches to evaluate the potential carcinogenicity and mutagenicity of pesticides and other agents. Authoritative groups, such as IARC,1 the U.S. Environmental Protection Agency (EPA), and the National Toxicology Program (NTP), have adopted a weight-of-theevidence approach to make decisions on the carcinogenicity of an agent. For example, after reviewing the human, animal, and relevant biological data for one class of pesticides, IARC concluded that the spraying and application of nonarsenical insecticides entail exposures that are probably carcinogenic to humans (IARC, 1991). To date, a relatively small number of pesticides (10) have been recognized by one or more of these organizations as human carcinogens (Goldman, 1998; IARC, 2009; NTP, 2000). It should be noted that in these cases, the primary evidence has come not from agricultural uses but, rather, from studies of exposed workers manufacturing the agent for other industrial uses or, in the case of inorganic arsenic, from therapeutic and industrial uses as well as environmental exposures (IARC, 1987a). For instance, 1

IARC, part of the World Health Organization, produces authoritative evaluations of the carcinogenic risks of chemicals and other agents to humans. Following a critical review of both human and animal studies, IARC classifies agents as exhibiting sufficient evidence of carcinogenicity, limited evidence of carcinogenicity, inadequate evidence of carcinogenicity, or evidence suggesting a lack of carcinogenicity in animals or humans. As a final step, IARC considers the entire body of evidence including mechanistic information to reach an overall evaluation of the carcinogenicity of the agent to humans. Other regulatory agencies, such as the U.S. EPA and the NTP, use similar approaches to evaluate the carcinogenicity of chemical agents.

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most of the evidence for the carcinogenicity of agents such as inorganic arsenic, benzene, cadmium, and chromium (VI), which historically were used as pesticides,2 as well as agents currently registered for use, such as ethylene oxide and coal tar creosote, has been obtained from studies involving nonagricultural uses. In some cases, it is believed that the agent responsible for the toxic effects seen in the pesticide-exposed individuals is a contaminant or an “inert” ingredient in the pesticide formulation rather than being the active ingredient itself. For example, many of the adverse effects proposed as being associated with the chlorophenoxyacetic acid herbicides are believed to be due to contamination by low levels of 2,3,7,8-tetrachlorodibenzo-para-dioxin (TCDD), a potent animal and human carcinogen (IARC, 1997; IOM, 1999, 2000). Furthermore, it is conceivable that other cancers such as the leukemias and non-Hodgkin’s lymphomas that have been attributed to pesticide exposure may in part be due to the use of benzene and other solvents as ingredients in the formulation products (Blair and Zahm, 1995; Petrelli et al., 1993). As indicated previously, chronic testing in animals is also used by regulatory agencies to evaluate the carcinogenic effects of chemical agents. Animal bioassays have been conducted for a considerable number of individual pesticides, and a significant portion of these have been reported to be tumorigenic in one or more animal tissues. According to Zahm and Ward (1998), of the 51 pesticides evaluated prior to 1990 by the U.S. National Cancer Institute and NTP, 24 exhibited carcinogenicity in chronic animal bioassays. These authors further reported that as of 1997, the IARC had classified 26 pesticides as having sufficient evidence of carcinogenicity in animals and 19 as having limited evidence. However, because of their cost, lengthy duration, and concern about relevance to humans, these bioassay results are considered as less than ideal and are usually evaluated in conjunction with additional types of biological information. In addition to human and chronic animal studies, regulatory agencies often rely on other relevant biological data to assist in the evaluation of carcinogenicity. These other data may include information on preneoplastic lesions, tumor pathology, genetic and related effects, structure– activity relationships, metabolism and pharmacokinetics, physicochemical parameters, and mechanisms of action (IARC, 1999e). In particular, short-term tests evaluating the genetic toxicity of the agent are often relied on in the decision-making process. The development and interpretation of these short-term tests have stimulated the development of the field of genetic toxicology.

2

Some forms of arsenic are still registered in the United States for use under severely restricted conditions. Arsenicals continue to be used as insecticides and wood preservatives in other countries (Zahm et al., 1997).

Hayes’ Handbook of Pesticide Toxicology

As a subspecialty of toxicology, genetic toxicology is concerned with the adverse effects of chemicals and other physical agents on the DNA and other genetic components of living organisms. The primary focus of this discipline is to identify the agents and mechanisms involved in the formation of mutations – heritable genetic alterations in cells. When broadly defined, mutagenesis encompasses the induction of DNA damage as well as all types of genetic alterations, ranging from a single nucleotide change in the DNA sequence to large-scale changes in chromosome structure and number. The recognition that cancer is fundamentally a genetic disease, combined with the close association that has been seen between mutagenicity and carcinogenicity, has led to the use of mutation and genotoxicity assays as screens to identify agents likely to be carcinogenic or cause other genetic diseases. During the past 40 years, a large number of short-term tests have been developed as screening tools to identify genotoxic and mutagenic chemicals. These short-term tests may employ bacteria, yeast, plants, insects, isolated mammalian cells, or whole animals and can be performed for a fraction of the cost and time required for a long-term cancer bioassay. In addition, a number of these assays have been modified for use in biomonitoring to detect genetic alterations occurring in human populations exposed to genotoxic and carcinogenic agents. The objective of this chapter is to provide an overview of methods of genotoxicity testing and their application to identifying pesticides capable of inducing genetic damage. The next section focuses on the most common short-term tests that are employed for detecting the genotoxicity of pesticides in model systems and the use of these assays to detect genetic alterations in exposed humans. This is followed by an overview of the results of genotoxicity studies that have been performed on individual agents and studies of genetic damage in pesticide-exposed workers. The last section briefly addresses the value and interpretation of this information in the safety evaluation and risk assessment process.

11.2  Genotoxicity tests Hundreds of short-term tests have been developed to screen chemicals for potential mutagenic and carcinogenic effects. These assays measure effects ranging from DNA adduct formation to mutations induced in transgenic animals. A listing of representative short-term tests as well as a brief description of how these effects are measured is presented in Tables 11.1 and 11.2. Each of these genotoxicity assays has its own unique characteristics and measures only a subset of the possible heritable alterations involved in cancer and other genetic diseases. As a result, combinations of short-term tests are often used to increase the likelihood of detecting genotoxic effects.

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Table 11.1  Representative Short-Term Tests for Genotoxicity Type of test

Specific test

DNA adduct formation

Direct measurement of covalently bound adducts Measurement of oxygen radical-derived adducts Covalent binding of radiolabeled chemicals 32 P-postlabeling of adducts Immunological detection of adducts

DNA damage in microorganisms

Pol A test rec test Mitotic recombination, mitotic crossing over, or mitotic gene conversion in yeast (D3, D4, D5, or D7 assays)

DNA damage in mammalian cells

Unscheduled DNA synthesis (UDS) Single-cell gel electrophoresis (Comet) assay Sister chromatid exchange (SCE)

Gene mutation in bacteria and fungi

Salmonella microsome reversion assay (Ames test) E. coli WP2 assay Yeast “forward” and “reverse” mutation assays Miscellaneous

Gene mutation in higher systems

HPRT, TK, and Na/K-ATPase assays in vitro Sex-linked recessive lethal assay in Drosophila Tradescantia or maize waxy locus plant tests HPRT assay in cells or in vivo Mutation in lac I/lac Z-bearing transgenic animals

Chromosomal effects in isolated cell systems

In vitro cytogenetics assays In vitro micronucleus test Aneuploidy assays

Chromosomal effects in whole organisms

In vivo cytogenetic assay in rodent bone marrow cells Mammalian erythrocyte micronucleus test Nondisjunction assay Heritable translocation assay Dominant lethal assay Alterations in germ cells

Oncogenic transformation

Transformation assays (clonal or focus)

Modified from U.S. EPA (1979).

Throughout the years, requirements for genotoxicity testing have been established in the United States and other developed nations, and agents being proposed as new pesticides must undergo testing prior to registration. The testing scheme used by the U.S. EPA is shown in Table 11.3. Using the U.S. EPA requirements as an example, the initial tier 1 test battery includes (1) a gene mutation assay involving at least five strains of bacteria, typically the Salmonella typhimurium reverse mutation assay and the WP2 test in Escherichia coli; (2) one of several gene-inactivating (forward) mutation assays using mammalian cells in culture (or a gene mutation assay combined with an in vitro cytogenetics assay); and (3) an in vivo assay for chromosomal effects in mammalian bone marrow cells using either metaphase analysis for structural aberrations or the micronucleus assay (Auletta et al., 1993; Dearfield et al., 1991). Depending on the results of the initial battery (as well as

other relevant information), additional tier 2 or tier 3 testing may be required to assess the pesticide’s potential to cause heritable mutations in germ cells or a chronic animal bioassay may be mandated to evaluate carcinogenic risks. As a general principle, agencies such as the U.S. EPA place greater weight on tests conducted in eukaryotes than in prokaryotes and in mammalian species rather than in submammalian species when conducting a hazard evaluation of a chemical (Auletta et al., 1993). For heritable noncancer risks, the results from studies in germ cells are accorded more weight than those obtained using somatic cells. Because of their prominent role in the testing of pesticides, the principal required assays are described in more detail later. For more detailed reviews of these and other short-term genotoxicity tests, the reader is referred to more comprehensive sources [IARC, 1980; International Programme of Chemical Safety (IPCS), 1985; Rice et al., 1999b].

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Table 11.2  The Measurement of Genotoxic Effects in Short-Term Tests DNA binding (32P-postlabeling, 8-OH-dG, and others) The covalent binding of a chemical to DNA is used as a measure of its reactivity and potential for genotoxicity. The DNA adducts can be detected and quantitated either directly or by using a radiolabeled chemical, by labeling the adducted nucleotide after formation, or by immunological techniques.

DNA damage in bacteria (Pol A test and rec test) Two strains of bacteria are used that are identical except in their ability to repair DNA damage; one strain can repair damage, whereas the other cannot. Both strains are exposed to the test substance, and the extent to which cells are killed is measured for each. If the repairdeficient strain has a greater degree of cell killing, DNA damage is assumed to have occurred.

DNA damage in yeast (mitotic recombination, mitotic gene conversion, or mitotic crossing over) Special strains of yeast cells are used to test for these effects. When the cells change color, typically from white to either pink or red, DNA damaging potential is indicated.

Chromosomal effects in isolated cells or whole organisms (cytogenetics assays or in vitro micronucleus assays) Treated cells (or cells from treated organisms) are stained and then examined under the microscope for various chromosomal abnormalities. Lost, broken, or misarranged chromosomes or the formation of micronuclei indicate genotoxicity.

Gene mutation in bacteria or fungi (Ames test, WP2 assay, yeast assays, and others) Special strains of bacteria are used that cannot grow without a nutritional supplement. Certain types of mutations will permit these bacteria to grow in unsupplemented media. By treating the cells and then seeing if they can grow in unsupplemented media, mutagenicity can be measured. Distinguishing mutated bacteria from nonmutated bacteria is not necessary using this procedure because only mutant cells can grow and form visible colonies.

Oncogenic transformation (transformation assays) When certain types of mammalian cells are treated in vitro with carcinogens, they undergo cancer-like transformation. If these cells are injected into appropriate experimental animals, tumors will appear. Most frequently, transformed cells are distinguished by their unusual growth patterns in culture, such as abnormal piling-up and disorientation of cells.

DNA damage in mammalian cells (unscheduled DNA synthesis, sister chromatid exchange, and single-cell gel electrophoresis “Comet” assay) Abnormal distribution of a DNA marker indicates whether DNA damage has occurred. Microscopic examination, photographic measurements, and computerized image analysis systems are used to detect the DNA damage.

Mammalian erythrocyte micronucleus test in vivo Animals are treated with a chemical, and their red blood cells are removed, stained, and examined under the microscope. If small nuclei containing genetic material (micronuclei) are observed, chromosomal damage has occurred. Normal red blood cells will not contain micronuclei.

Gene mutation in mammalian cells or plants Mammalian cells (HPRT, TK, and Na/K-ATPase assays): In these systems, mutations that confer resistance to a poison are measured. Cells are first treated with a test chemical and then exposed to the poison. Because only mutant cells can survive and grow, mutagenicity can be measured simply by observing the extent of growth in the poisonous environment. Plant cells (Tradescantia and maize waxy locus): Mutations in these plants are detected by looking for color changes in the stamen hairs (Tradescantia) or pollen grains (maize).

Drosophila melanogaster (sex-linked recessive lethal test for mutations; nondisjunction and heritable translocation assays for chromosomal effects) Drosophila have a variety of “marker” traits that can be used to signal whether gene mutations or chromosome disturbances have occurred. Specially “marked” male or female flies are treated with a substance, mated, and then their offspring are observed to see if they have certain specific features, such as unusual eye color or shape.

Modified from U.S. EPA (1979).

11.2.1  Bacterial Reverse Mutation Assay The bacterial reverse mutation assay uses specially engineered amino acid–requiring strains of S. typhimurium and E. coli to detect point mutations, which involve the substitution, deletion, or insertion of one or a few DNA base pairs (IPCS, 1985; U.S. EPA, 1998c). The widely used Salmonella assay was developed by Ames, McCann, and Yamasaki and is commonly known as the Ames test (Ames et al., 1975; McCann and Ames, 1976). The basis for the assay is as follows: Following exposure to a mutagenic chemical, mutations are detected that reverse existing

gene-inactivating mutations present in the Salmonella test strains, thereby restoring the ability of the bacteria to synthesize the essential amino acid. The bacteria carrying the reverse mutations (called revertants) are detected by their ability to grow in the absence of the amino acid required by the parental test strain. Many of the test strains have also been engineered to increase the sensitivity of the assay. These enhancements include a modification of the cell wall to be more permeable to lipophilic chemicals, inactivation of a gene involved in DNA excision repair, and addition of another gene coding for an error-prone DNA repair gene.

Chapter | 11  Genotoxicity of Pesticides

Table 11.3  Genotoxicity Testing Requirements for Pesticides Registered in the United States Tier 1 (in vitro and in vivo tests) A bacterial reverse mutation assay involving at least five test strains In vitro gene mutation assay (or a gene mutation test combined with in vitro cytogenetics) In vivo cytogenetic assay (structural chromosome aberrations or micronuclei) Tier 2 (in vivo tests) Tests detecting test agent interaction with gonadal DNA Dominant lethal assay Tier 3 (in vivo tests) Specific locus assay (biochemical or visible) Heritable translocation test Adapted from Dearfield et al. (2002) and Cimino (2006).

In addition, the assay is conducted in the presence and absence of a mammalian metabolic system to increase the sensitivity of the assay to chemicals requiring metabolic activation for genotoxicity. Most commonly, the metabolic system is a cofactor-supplemented post-mitochondrial fraction (S9) prepared from the livers of rats treated with enzyme-inducing agents such as Aroclor 1254. By using these specially engineered and rapidly growing bacteria, chemically induced mutations occurring at low frequencies (1  106) in tens to hundreds of millions of bacteria can be rapidly and inexpensively detected. However, the targets for reverse mutations in the test strains are very small in relationship to the bacterial genome, and only a narrow range of point mutations can be detected. To increase the range of point mutations that can be detected, regulatory guidelines recommend that five different test strains of bacteria be used in the presence and absence of the metabolic activation system (U.S. EPA, 1998c). This assay has been shown to be from moderately to highly efficient at predicting carcinogenicity with predictive values generally ranging from 50 to 90% depending on the study and the characteristics of the chemicals being tested (Brusick, 1987). However, due to fundamental differences between prokaryotic and eukaryotic organisms, this assay is not able to detect mutations induced by some types of chemical agents, such as topoisomerase inhibitors and nucleoside analogues (U.S. EPA, 1998c).

11.2.2  In Vitro Mutation Assay   in Mammalian Cells The in vitro mammalian cell gene mutation assay can be used to detect gene-inactivating mutations induced by chemicals (U.S. EPA, 1998d). Mouse, Chinese hamster,

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or human cell lines are exposed to the test chemical and mutations occurring in endogenous genes such as thymidine kinase (TK), hypoxanthine-guanine phosphoribosyl transferase (HPRT), and a transgene of xanthine-guanine phosphoribosyl transferase (XPRT) are measured. Using the assay for mutations in thymidine kinase as an example, cells with a mutation converting the TK heterozygote (TK/) to cells lacking a functional TK allele (TK/) are resistant to the cytotoxic effects of the nucleotide analogue trifluorothymidine (TFT). Thymidine-proficient cells are sensitive to TFT, which inhibits cellular metabolism and halts cell division. As a result, mutant cells are able to proliferate in the presence of TFT, whereas normal cells that contain the functional TK allele are unable to grow. Similarly, cells deficient in HPRT or XPRT are selected based on their resistance to 6-thioguanine or 8-azaguanine, respectively. In these assays, cells are exposed to the test chemical in both the presence and the absence of the S9 metabolic activation for a suitable period of time and then subcultured to allow phenotypic expression prior to mutant selection with the toxic nucleotide analogue. Mutant frequency is then determined, after an appropriate incubation period, by seeding known numbers of cells in medium containing the selection agent to detect mutant cells and in medium without the selection agent to determine the cloning efficiency. The principal advantage of this assay is that it allows rare mutations occurring in mammalian cells to be detected simply and relatively inexpensively. Moreover, because this assay measures gene-inactivating mutations in eukaryotic cells, it is capable of detecting a much broader range of mutagenic events (i.e., large deletions, recombination, etc.) than the bacterial mutation assays.

11.2.3  In Vivo Cytogenetic Assay The in vivo chromosome aberration assay is used for the detection of structural chromosome aberrations induced by test chemicals in the bone marrow of mammals, typically rodents (U.S. EPA, 1998e). In this assay, animals are administered the test substance by an appropriate route of exposure and are sacrificed at selected times (typically 12–36 h) after treatment. Prior to sacrifice, the animals are treated with a spindle-disrupting agent to arrest rapidly dividing bone marrow cells in the metaphase stage of the cell cycle. Chromosome preparations are made from the bone marrow cells and, following staining, the metaphase cells are analyzed for structural damage to the chromosomes. Some information on changes in chromosome number (aneuploidy and polyploidy) can also be obtained. Because a chromosome break can occur within most, if not all, DNA sequences throughout the genome, this assay is believed to be highly sensitive at detecting agents inducing double-stranded breaks in DNA. In previous studies, it has been shown that when tested, most human cancer-causing agents induce increased levels of chromosome aberrations

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in the bone marrow of rodents (Ashby and Paton, 1993). Moreover, this assay is thought to be particularly valuable in that chromosomal alterations are the underlying cause of many genetic diseases and play an important role in carcinogenesis. The in vivo aberration assay is considered particularly useful for assessing mutagenic hazards in that it allows normal in vivo metabolism, toxicokinetics (absorption, distribution, and excretion), and DNA repair processes to occur (Auletta et al., 1993).

11.2.4  Micronucleus Assay The micronucleus assay is similar to the in vivo aberration assay in that both measure chromosome alterations in treated mammals and, according to most regulatory guidelines, either can be used in the initial testing (Auletta et al., 1993; Dearfield et al., 1991). The micronucleus assay detects chromosome breakage and loss occurring following chemical treatment. Although micronuclei can be formed in any dividing tissue of any species following treatment, for regulatory purposes the assay is almost always conducted in the bone marrow or, less frequently, the peripheral blood erythrocytes of rodents (U.S. EPA, 1998f). As a bone marrow erythroblast develops into a newly formed RNA-containing (polychromatic) erythrocyte, the main nucleus is extruded. In a damaged cell, the micronucleus that has been formed remains behind in the anucleate cytoplasm. Using a stain such as acridine orange that differentially stains RNA and DNA, the DNA-containing micronucleus can easily be visualized in the cytoplasm of the newly formed RNA-containing erythrocytes. An increase in the frequency of micronuclei following treatment with a test chemical indicates that an increase in chromosome damage has occurred. The assay can be performed in one of two ways: with a single dose followed by two or more sampling times or with two or more sequential doses followed by a single harvest. As with the in vivo aberration assay, this in vivo assay allows normal metabolism, toxicokinetics, and DNA repair to occur. In addition, many human and animal carcinogens when tested have shown positive results in this assay (Ashby and Paton, 1993).

11.3  Genotoxicity testing   of pesticides As indicated previously, genotoxicity testing is required for the registration of new pesticides in the United States and most developed nations. Testing has also been performed for many of the pesticides that were registered prior to the current testing requirements. It should be noted, however, that often the results of these tests have been considered proprietary and have not been published in the public domain. Published genotoxicity test results for many pesticides and

Hayes’ Handbook of Pesticide Toxicology

other agents evaluated by the U.S. EPA and IARC are available in the Genetic Activity Profiles (GAP) database (Waters et al., 1991, 1999) or in the IARC monograph series or its supplements (for examples, see IARC, 1987c,d, 2008). Two valuable sources of the summary results of unpublished tests on pesticides are the toxicological summaries compiled by the California Department of Pesticide Regulations (available at www.cdpr.ca.gov/ docs/toxsums/toxsumlist.htm) and the toxicological evaluations performed as part of the joint meeting of the Food and Agricultural Organization panel of experts on pesticide residues in food and the environment [for an example, see Food and Agricultural Organization/World Health Organization (FAO/WHO), 1999]. A representative listing of specific pesticides, their activity in various genotoxicity tests, and evaluations for carcinogenicity is shown in Table 11.4. As is evident from the table, there are a variety of patterns of responses. Some agents are clearly genotoxic and carcinogenic, whereas others have shown activity in the genotoxicity assays without showing an increase in tumors in the cancer bioassays. Other pesticides have primarily exhibited negative results in short-term genotoxicity assays but have shown increases in tumors in chronic animal testing. Some agents have demonstrated no genotoxic or carcinogenic effects in in vitro or in vivo studies. Finally, many agents have given mixed or equivocal responses in genotoxicity or carcinogenicity tests. Interpretation of this latter pattern of responses is particularly challenging due to the likelihood of false positive results when many short-term assays are conducted or assays are performed under conditions (high concentrations, increased osmolality, pH, oxygen tension, etc.) that may differ significantly from those likely to be encountered in vivo. In addition, the pathological evaluation of many different tissues and organs also increases the likelihood of false positives in a chronic animal cancer bioassay. For illustration, examples of each of the preceding patterns of response are presented.

11.4  Patterns of response 11.4.1  Pesticides Exhibiting Both Genotoxicity and Carcinogenicity 11.4.1.1  Ethylene Oxide Ethylene oxide or epoxyethane is an insecticidal fumigant used for stored food products, bedding, carpets, and clothing (Gehring et al., 1991). It is also used to sterilize heatsensitive medical devices and as an intermediate in the synthesis of other chemicals, particularly ethylene glycol (Dellarco et al., 1990). Structurally, it is a reactive chemical that exerts its cytotoxic effects by alkylating a broad range of critical cellular macromolecules such as DNA and

Chapter | 11  Genotoxicity of Pesticides

363

Table 11.4  Short-Term Genotoxicity Results and Evaluation of Carcinogenic Risk for Selected Pesticides Pesticide

Mutation

Chromosomal aberrations/ micronuclei

IARC classification

Salmonella (Ames test)

Mammalian cells

In vitro

In vivo

Human carcinogenicity

Animal carcinogenicity





a

1a, 1b

SE

LE



a

h

SE

SE

a

ah

Inorganic metals Arsenic compounds Cadmium chloride Chromium VI compounds



b

e ,e

 , 1

a

b

a





,

 ,

SE

SE





a

a, eb

NR

NR

a

ab

Carbamates Propoxur*

h

Carbaryl





 , 1



ND

IE

Aldicarb



,1h

1ah

1a

ND

IE

Chlorinated hydrocarbon insecticides Chlordane





0

0

IE

SE

Heptachlor



1

0

0

IE

SE

DDT





ea, -1ah

a, ()ah

IE

SE

Aldrin



0

1ah

1b, 1a

IE

LE

Endosulfan*









NR

NR

Endrin



()

a

0

IE

IE

Deltamethrin





0

a, b

ND

IE

Fenvalerate



0

1ah

a, 1b

ND

IE

1

a

NR

NR

a

ND

IE

Pyrethroids

Cypermethrin Permethrin

 

1

0



0

1

1

a, 1ah

a

Organophosphate insecticides Dichlorvos



Parathion





0

Methyl parathion







Malathion



0



Diazinon

 *

e

 a

IE

SE

a

ND

IE

a

ND

ESL

ND

IE

NR

NR

NR

NR

 e

a

 , 1

1

b

b

Chlorpyrifos



1

1



Isazofos



0

0



NR

NR

Captafol



0

a, bh, 1ah, 1b

0

ND

SE

Pentachlorophenol





ea, 1ah

1ah

IE

SE

e

a

b

IE

IE

ND

LE

Fungicides

Thiram Ziram

 

0

1

a

 , 1 ah

1 , 1

a

a b

ah

1 , 1 , 1

(Continued )

Table 11.4  (Continued) Pesticide

Mutation Salmonella (Ames test)

Mammalian cells

ortho-Phenylphenol*



1

Chlorothalonil



Hexachlorobenzene* 1,4-Dichlorobenzene* Propylene oxide

Chromosomal aberrations/ micronuclei In vitro

IARC classification

In vivo

Human carcinogenicity

Animal carcinogenicity

ea

, 1b

IE

LE



ea, ah

a, b

IE

SE



0

1a, 1ah, 1b, 0 1bh

IE

SE





a, 1b, -1bh

a, b

IE

SE

b

IE

SE

IE

SE

 , 1

ND

LE

a



1

1 , 1



1

a

ah

1

Herbicides Atrazine

1b a

b

a

Monuron



E

1

Picloram



0

1ah

1a

ND

LE

Simazine



1

0

1b

IE

LE

Trifluralin





eah

a

IE

LE

MCPA





1

1a, b, ah

IE

ND

2,4-D



0

ah

ah, b

LE

IE

Methyl chloride



1

0

0

IE

IE

Bentazon*



1

0

1b

NR

NR

0

b

LE

IE

IE

SE

IE

SE

LE

SE

IE

SE

2,4,5-T



Amitrole



0

a

1 , 1 a

b

e

1



a, 1b, 1ah



Fumigants and nematocides Acrylonitrile

 *

Ethylene oxide

 *

Ethylene dibromide Formaldehyde

 

*

Methyl bromide DBCP

Carbon tetrachloride

  

h

a

1



a

b

a

ah

 e 

 , 1

b

ah

a

SE

SE

b

a

IE

LE

b

as

IE

SE

a

b

 , 1

IE

SE

ah

 , 1

a

 , 1 ah

b

bhf

 ,, ,

a



a

1 , 

, e

b

,





a, 1b

b

1 , 1

Tetrachloroethylene



1

1�

1

LE

SE

1,3-Dichloropropene



e





ND

SE

Xylene





1a, -1ah

b

IE

IE

Benzene





a

a, b

SE

SE

Piperonyl butoxide





0

0

ND

IE

Solvents and others

a, chromosomal aberrations; b, micronucleus; h, human cells; s, spermatogonia; (), weakly positive; e, equivocal/inconclusive; 0, no test results were located; 1, positive in one study; 1, negative in one study; , positive in more than one study or the majority of studies; , negative in more than one study or the majority of studies; f, micronucleus formation was positive in buccal mucosal cells in humans, whereas it was negative in peripheral blood lymphocytes, possibly due to the high reactivity of formaldehyde at the primary site of exposure; IE, inadequate evidence for carcinogenicity; LE, limited evidence for carcinogenicity; SE, sufficient evidence for carcinogenicity; ESL, evidence suggesting lack of carcinogenicity; ND, no adequate data were available; NR, not reviewed by IARC. *See text for additional details. This table was compiled primarily from five sources: (1) The IARC Monographs on the Evaluation of Carcinogenic Risks in Humans, (2) the Environmental Health Criteria series published by the International Programme on Chemical Safety, (3) the toxicological evaluations performed as part of the joint meeting of the FAO panel of experts on pesticide residues in food and the environment, (4) the Genetic Activity Profile database generated jointly by the U.S. EPA and IARC, and (5) the toxicological data review summaries prepared by the California Department of Pesticide Regulation.

Chapter | 11  Genotoxicity of Pesticides

proteins (Dellarco et al., 1990; Gehring et al., 1991). Given its ability to alkylate DNA, it is not surprising that it exhibits genotoxic effects in most genotoxicity assays. Ethylene oxide has been shown to be mutagenic in bacterial and mammalian cells, increase chromosome aberrations and micronuclei in the bone marrow of rodents, and exhibit positive responses in a series of other genotoxicity assays (Dellarco et al., 1990; IARC, 1994). In reviewing the evidence, IARC has concluded that ethylene oxide is both an animal and a human carcinogen (IARC, 1994, 2008). In addition to affecting somatic cells, ethylene oxide is also an established germ cell mutagen that has been shown to induce dominant-lethal mutations and translocations in rodents (Dellarco et al., 1990). Ethylene oxide is one of the few agents for which heritable risks to humans have been evaluated (Rhomberg et al., 1990).

11.4.1.2  Ethylene Dibromide Ethylene dibromide, EDB or 1,2-dibromoethane, has been used as a fumigant for stored grain, fruits, and vegetables (Gehring et al., 1991). It has also been used as a soil treatment for nematodes and as a scavenger in tetraethyl leadcontaining gasoline. EDB is metabolically activated through both microsomal- and glutathione transferase-dependent pathways to form reactive DNA and protein-binding metabolites (Gehring, et al., 1991). EDB has been shown to be mutagenic in bacteria and mammalian cells, to bind to DNA, to induce DNA strand breakage, and to increase unscheduled DNA synthesis (U.S. EPA, 1997). Although genotoxic in the majority of in vitro tests and in vivo assays for DNA breakage, EDB has shown largely negative results in in vivo assays of chromosome damage and dominant lethal mutations (IARC, 1999b). These somewhat differing results may reflect the target organ specificity, as well as the specific types of DNA damage induced by this agent. EBD has been shown to exhibit carcinogenic effects in multiple animal species. However, the evidence for carcinogenic effects in humans is considered inadequate (IARC, 1999b; U.S. EPA, 1997). Both IARC and the U.S. EPA consider EDB to be a probable human carcinogen. Similar patterns can be seen for other pesticides or “inert” ingredients such as chromium IV, arsenic, formaldehyde, benzene, and coal tar creosote. In each of these cases, the agent is carcinogenic in animals and/or humans, and it is positive in most genetic toxicity assays.3 The lack of activity in a few short-term tests suggests that the agent acts through a specific genotoxic mechanism, that target organ-specific effects or metabolism may be occurring, or that the genotoxicity result is in error (i.e., a false negative). Based on 3

Benzene and arsenic have exhibited negative results in gene mutation assays but have been positive for the induction of chromosomal alterations in vivo. The critical genetic alterations in the carcinogenicity of these agents appear to be chromosomal in nature.

365

the strongly positive results observed, most of these types of agent have been banned for use as pesticides or are registered for use under highly restricted conditions.

11.4.2  Pesticides Exhibiting Genotoxicity with Limited or No Evidence of Carcinogenicity 11.4.2.1  Methyl Bromide Methyl bromide or bromomethane has been widely used as a fumigant for control of insects, nematodes, fungi, and weeds (Gehring et al., 1991; IPCS, 1995). Although methyl bromide has been shown to react with both DNA and proteins, its mechanism for toxicity remains to be elucidated (IPCS, 1995). Methyl bromide has been shown to be genotoxic in most short-term genotoxicity tests (IARC, 1999c; IPCS, 1995): It induced mutations in bacteria and mammalian cells, increased the incidence of micronuclei in vivo in mouse and rat bone marrow erythrocytes, and was shown to bind covalently to the DNA in several rat and mouse organs. In contrast, methyl bromide has produced mixed, largely negative responses in chronic animal bioassays. In a short 13-week study in which methyl bromide was administered by oral gavage, it was reported to produce squamous cell carcinomas of the forestomach (IARC, 1999c; U.S. EPA, 1990). However, this result was questioned, and upon reexamination of histological slides, a group of NTP pathologists concluded that the lesions were hyperplasia and inflammation rather than neoplasia (U.S. EPA, 1990). In inhalation studies, the most relevant route of human exposure, methyl bromide has been reported to be largely negative, although there is limited evidence for tumorigenicity in various tissues. According to IARC, no significant increase in tumors was observed in two inhalation studies in mice and one in rats (IARC, 1999c). In another rat inhalation study, a significant increase in pituitary gland adenomas was seen in males treated at the highest dose. However, the conclusions of these studies continue to be controversial because a detailed examination of two of the inhalation studies previously described as negative led some reviewers to suggest that methyl bromide was capable of inducing tumors in some tissues (CDPR, 1999b). Based on its evaluation of the literature, IARC has concluded that there is limited evidence for the carcinogenicity of this agent in experimental animals (IARC, 1999c), whereas the U.S. EPA has considered the data inadequate to reach any conclusion (U.S. EPA, 1990). Both IARC and the U.S. EPA have stated that there is inadequate evidence to make conclusions about the carcinogenicity of methyl bromide in humans (IARC, 1999c; U.S. EPA, 1990). As indicated previously, methyl bromide is genotoxic in most in vitro and in vivo assays. Although there is some evidence for the carcinogenicity of methyl bromide, it

366

has not exhibited consistent carcinogenic effects in most studies. The basis for the discrepancy between the shortterm tests and the animal bioassay results is not clear. The negative cancer bioassay results could be false negatives reflecting inadequacies of the animal tests. However, several bioassays have been conducted with similar results, and no carcinogenic effects were seen even in the comprehensive mouse bioassay conducted by the NTP (1992). Alternatively, methyl bromide may alkylate DNA in vivo at sites that are readily repaired or lead directly to cell lethality rather than heritable mutations.

11.4.3  Pesticides Exhibiting Carcinogenicity without Appreciable Genotoxicity 11.4.3.1  Propoxur Propoxur, or Baygon, is an important carbamate insecticide used primarily against household insects and pests of domestic animals. It has been considered among the top 10 most widely used home and garden pesticides in the United States (Grossman, 1995). Similar to other carbamate insecticides, propoxur inhibits acetylcholinesterase, an enzyme involved in neurotransmission, producing neurotoxic effects in insects and nontarget organisms. Propoxur has yielded negative results in the majority of short-term genotoxicity tests that have been conducted (FAO/WHO, 1990). It was negative in bacterial and mammalian mutation assays and in bacterial DNA repair assays. It was reported to be negative in most chromosome aberration and micronucleus assays in vitro and in vivo (FAO/WHO, 1990), although at least two positive studies have been published (Agrawal and Mehrotra, 1997; Wei et al., 1997). In chronic studies, no evidence of carcinogenic effects was seen in mice treated with propoxur for 24 months or hamsters treated for 53 weeks (FAO/WHO, 1990). However, in a series of studies conducted in rats, highly significant increases in hyperplasia and bladder tumors were seen at high doses of propoxur. No hyperplastic effects in the bladder were seen in short-term studies employing mice, dogs, or monkeys, whereas effects were seen in short-term studies in the bladders of Sprague– Dawley rats. Dietary studies have indicated that in addition to high doses, the urothelial effects of propoxur are dependent on high urinary pH. The carcinogenic effects of propoxur in rats have been proposed to be due to chronic mitogenic stimulation of propoxur or a metabolite on the urothelium rather than from a direct genotoxic or mutagenic effect (Cohen et al., 1994).

11.4.3.2  Hexachlorobenzene Historically, hexachlorobenzene (HCB) was commonly used as a seed treatment for prevention of fungal growth

Hayes’ Handbook of Pesticide Toxicology

on crops such as wheat, barley, oats, and rye (IPCS, 1997). Concern for human health and the environment resulted in its discontinued use as a pesticide in many countries during the 1970s. HCB is currently found as an unintentional byproduct in several high-volume chlorinated solvents (carbon tetrachloride, trichloroethylene, and perchloroethylene) and in various pesticides, including pentachloronitrobenzene, chlorothalonil, DCPA, picloram, and pentachlorophenol (Agency for Toxic Substances and Disease Registry, 1997). In general, studies investigating the genotoxicity of HCB have indicated that it exhibits weak or no genotoxic activity (Brusick, 1986; Gorski et al., 1986; IPCS, 1997). In most studies, HCB exhibited no detectable mutagenic activity in Salmonella either with or without microsomal activation. No increase in structural chromosome aberrations was seen in Chinese hamster lung cells (Ishidate et al., 1988). Canonero and associates (1997) evaluated the in vitro genotoxicity of HCB in primary cultures of rat and human hepatocytes. An induction of micronuclei but not DNA strand breaks was seen in rat hepatocytes treated with HCB. In the studies with human hepatocytes, the authors reported that HCB induced a weak but significant increase in the frequency of both DNA breaks and micronuclei. Low levels of DNA binding were seen following the in vivo treatment of rats with HCB (Gopalaswamy and Nair, 1992). In addition, no increase in sister chromatid exchanges (SCEs) in the bone marrow of male mice or DNA fragmentation in the liver of rats was observed in HCB-treated animals (Gorski et al., 1986). HCB also failed to induce dominant lethal mutations in male rats (Simon et al., 1979). In contrast with the largely negative results in the genotoxicity studies, HCB exhibited carcinogenic effects in a series of animal studies, increasing the incidence of tumors in rats, hamsters, and mice (IARC, 1987b; IPCS, 1997; U.S. EPA, 1985, 1996). Increased tumor formation was seen in the liver and kidney as well as the adrenal, parathyroid, and thyroid glands of the treated animals. To date, the mechanisms underlying carcinogenesis in these organs remain unclear. Several theories have been proposed to explain the basis for certain tumors induced by HCB. For example, it has been proposed that the liver tumors occur as a secondary effect resulting from chronic toxicity to this organ (Carthew and Smith, 1994). It has been postulated that the male kidney tumors were due to an accumulation of the male rat-specific protein alpha 2u-globulin in the proximal renal tubular cells, resulting in a sustained cell proliferation and eventually neoplasia in this organ (Bouthillier et al., 1991). Lastly, others have proposed that the thyroid tumors were the result of a chronic stimulation of cell proliferation in the thyroid gland due to a chronic imbalance in thyroid hormones resulting from an induction of glucuronosyl transferases by HCB (Deutsche Forschungsgemeinschaft, 1998). All of these theories indicate that HCB exerts its

Chapter | 11  Genotoxicity of Pesticides

carcinogenic effects through indirect or “nongenotoxic” mechanisms. Assuming that these mechanisms are correct, the difference in the observed genotoxicity and carcinogenicity results would be expected. Following a review of the data, IARC and the U.S. EPA determined that there was sufficient evidence to conclude that HCB induces cancer in laboratory animals (IARC, 1987b; U.S. EPA, 1996). The evidence in humans is inadequate to draw definite conclusions. However, for regulatory purposes, the U.S. EPA considers HCB to be a probable human carcinogen (U.S. EPA, 1996), whereas IARC considers it to be a possible human carcinogen (IARC, 1987b).

11.4.4  Nongenotoxic Agents without Evidence of Carcinogenicity 11.4.4.1  Endosulfan Endosulfan, or thiodan, is a chlorinated insecticide used on a wide variety of food and non-food crops, including grapes, cantaloupes, lettuce, tomatoes, alfalfa, and cotton. Although a few positive responses have been reported in short-term tests (Smith, 1991), endosulfan is generally viewed by regulatory bodies as being nongenotoxic (CDFA, 1988; FAO/WHO, 1999). Endosulfan has primarily exhibited negative results in both bacterial and mammalian cell gene mutation assays. It was also negative in inducing chromosome aberrations or micronuclei in vitro as well as in vivo. In addition, it has been reported to be negative in other genotoxicity assays. Endosulfan did not exhibit carcinogenic effects in chronic bioassays conducted using mice or rats (FAO/WHO, 1999). Epidemiological studies of cancer in humans have not been conducted.

11.4.4.2  Chlorpyrifos Chlorpyrifos, or Dursban, is a broad-spectrum organophosphate insecticide with widespread usage on food commodities, turf, and ornamental plants. It has been commonly used indoors and for structural pest control. It is one of the most widely used pesticides in the United States, and until recently, it was one of the top five insecticides used in residential settings (U.S. EPA, 1999). In common with other organophosphate insecticides, upon bioactivation chlorpyrifos inhibits acetylcholinesterase, an enzyme involved in neurotransmission, producing neurotoxic effects in insects and nontarget organisms. Consequently, genotoxic effects would not be expected nor are they seen (CDPR, 1999a; U.S. EPA, 1999). Chlorpyrifos did not induce gene mutations in either bacterial or mammalian systems, although it was reported to induce slight increases in genetic alterations in yeast as well as DNA damage in bacteria. No increase in chromosome aberrations was seen in an in vitro

367

study using rat lymphocytes or in two in vivo studies evaluating micronuclei in the mouse bone marrow. It was ineffective at inducing unscheduled DNA synthesis in isolated rat hepatocytes. Chlorpyrifos was evaluated for carcinogenic potential in both rats and mice with no evidence of carcinogenicity (CDPR, 1999a; U.S. EPA, 1999).

11.4.4.3  Bentazon Bentazon, 3-(1-methylethyl)-1H-2,1,3-benzothiadiazin4(3H)-one-2,2-dioxide, is a herbicide used in agriculture for control of broadleaf weeds in crops such as soybeans, rice, corn, peanuts, and lima beans (U.S. EPA, 1998g). As summarized from U.S. EPA reports (1998b,g), bentazon is not chemically reactive and no highly reactive species have been identified during its metabolism. Bentazon was negative in bacterial mutation assays, in a mammalian cell assay, in the unscheduled DNA synthesis assay, and in the mouse micronucleus assay in vivo. In chronic animal bioassays, no increases in tumors were seen in the rat. A slight dose-related increase in hepatocellular tumors was seen in the mouse studies. However, upon reexamination, it was concluded that the incidence did not differ significantly from the controls. In its evaluation of the toxicity of bentazon, the U.S. EPA concluded that bentazon was essentially noncarcinogenic in animals and was not likely to cause cancer in humans.

11.4.5  Pesticides Exhibiting Mixed Results   in Genotoxicity or Cancer Tests 11.4.5.1  ortho-Phenylphenol ortho-Phenylphenol (OPP) and its sodium salt, sodium o-phenylphenate (SOPP), are broad-spectrum fungicides and disinfectants with widespread agricultural, industrial, and domestic usage. OPP has historically been among the most widely used home and garden pesticides (Grossman, 1995). Investigations into the genotoxic effects of SOPP and OPP have indicated that these compounds are inactive or weakly active in bacterial mutation assays (NTP, 1986). Some evidence for the mutagenicity of OPP has been seen in mammalian cell assays. A weak increase in mutations was seen at the TK locus in treated Chinese hamster ovary (CHO) cells (NTP, 1986), whereas a strong increase in ouabain-resistant mutants was reported to occur in a ultraviolet-sensitive human Rsa cell line following treatment with OPP (Suzuki et al., 1985). Negative results were also observed when measuring unscheduled DNA synthesis in rat hepatocytes following exposure to SOPP (Reitz et al., 1983). In cytogenetic studies, several reports indicate that OPP and its metabolite phenylhydroquinone have induced SCEs and structural chromosomal aberrations in CHO cells in the presence of exogenous metabolic

368

activation (NTP, 1986; Tayama et al., 1989; Tayama and Nakagawa, 1991; Tayama-Nawai et al., 1984), whereas others have reported negative or ambiguous results (Ishidate, 1988; NTP, 1986). Phenylhydroquinone was also shown to induce chromosome-containing micronuclei upon prostaglandin[H]synthase-mediated activation in V79 cells (Lambert and Eastmond, 1994). Following the in vivo administration of radiolabeled OPP and SOPP to male F344 rats, no increases in the covalent binding of these compounds to rat bladder DNA were observed using either liquid scintillation counting (Reitz et al., 1983) or a highly sensitive accelerator mass spectrometric technique (Kwok and Eastmond, 1997). Binding to bladder proteins was seen in both studies. Contradictory results have been reported for DNA binding using the 32P postlabeling technique, with one group reporting detectable OPP-derived adducts (Ushiyama et al., 1992) and another, focusing on adduct formation in the target urothelial cells, reporting negative results (Smith et al., 1998). A modest increase in DNA breakage in the bladder was detected in rats (Morimoto et al., 1989) and mice (Sasaki et al., 1997) following treatment with OPP or SOPP. In addition, significant increases in micronucleated bladder cells have been reported in rats administered high doses of OPP in the diet (Balakrishnan and Eastmond, 2006; Balakrishnan et al., 2002). OPP and SOPP have been tested for carcinogenicity in both mice and rats by administration in the diet. Increases in bladder tumors have been seen in multiple rat studies following treatment with OPP and SOPP (CDPR, 1997; IARC, 1999d). SOPP appears to be more potent and consistent in inducing carcinogenic effects, and it has been proposed that urinary pH plays an important role in the bioactivation and carcinogenesis of these compounds (Fujii et al., 1987; Kwok and Eastmond, 1997). The effects appear to be specific to the rat because little evidence of carcinogenicity was observed in chronically treated mice (IARC, 1999d), and bladder toxicity was not seen in shortterm studies in mice, guinea pigs, hamsters, and dogs (Cosee et al., 1992; Hasegawa et al., 1990). Upon review of the data, IARC concluded that OPP was not classifiable as to its carcinogenicity to humans and that SOPP was possibly carcinogenic to humans (IARC, 1999d). The mechanisms underlying the carcinogenic effects of OPP remain to be fully elucidated. It has been proposed that OPP acts as a bladder carcinogen in rats by inducing cytotoxicity and hyperplasia without directly binding to DNA (Smith et al., 1998). As a result, the observed genotoxicity may be indirect, occurring through the formation of oxygen radicals, through an enhancement of spontaneous mutations, through DNA damage closely associated with cytotoxicity, or through an interaction with protein targets (Appel, 2000; Balakrishnan and Eastmond, 2006; Kwok and Eastmond, 1997). The inconsistent results seen

Hayes’ Handbook of Pesticide Toxicology

in the short-term tests may, in part, be a reflection of this indirect mechanism of genotoxicity.

11.4.5.2  1,4-Dichlorobenzene 1,4-Dichlorobenzene, or para-dichlorobenzene (p-DCB), is commonly used to control moths, molds, and mildew and as a bathroom deodorizer. p-DCB is also used as an intermediate in the synthesis of polyphenylene sulfide resin. The genotoxicity of p-DCB has been investigated with mixed, largely negative results (IARC, 1999a). p-DCB was not mutagenic in bacteria or mammalian cells in vitro but did exhibit some evidence of DNA damage and mutagenicity in yeast. p-DCB produced mixed results in in vitro cytogenetic assays with both positive and negative reports for micronuclei and SCEs. It was negative in inducing DNA strand breaks and chromosome aberrations in vitro. p-DCB failed to exhibit genotoxic effects in vivo, exhibiting negative responses in unscheduled DNA synthesis, in the chromosome aberration assay, in the dominant lethal assay, and in the in vivo micronucleus assay. It was reported as positive in one DNA strand breakage assay and in one in vivo micronucleus assay. p-DCB bound to DNA in the liver, lung, and kidney of mice but not in that of male rats (IARC, 1999a). It also induced DNA damage in liver and spleen but not in kidney, lung, or bone marrow of mice. IARC stated that no conclusion could be drawn from the few data on genotoxicity in vivo (IARC, 1999a). In contrast to the negative genotoxicity results, p-DCB induced carcinogenic effects in both rats and mice. Follo­ wing oral administration, p-DCB increased the incidence of liver tumors in male and female mice as well as the incidence of renal carcinomas in male rats (IARC, 1999a). In evaluating the significance of these tumors, IARC concluded that the evidence did not support a mechanism of renal cell tumor formation that involved a direct interaction between p-DCB or its metabolites with DNA. The male kidney tumors induced by p-DCB were due to an accumulation of the male rat-specific protein alpha 2u-globulin in the proximal renal tubular cells that eventually resulted in neoplasia in this organ. This mechanism is widely accepted as not being relevant to humans (IARC, 1999a; Rice et al., 1999a; U.S. EPA, 1991). However, the IARC Working Group had more concern for the liver tumors that were seen at a high incidence in the male and female mice. Because p-DCB was reported to cause DNA damage in the liver and spleen of mice and to bind weakly to DNA, the tumors in the liver were thought to be potentially relevant to humans. IARC concluded that p-DCB was an animal carcinogen and possibly carcinogenic to humans (IARC, 1999a). As illustrated in the preceeding examples, different patterns of genotoxic and carcinogenic effects can be seen in short-term tests and in animal bioassays. In many

Chapter | 11  Genotoxicity of Pesticides

instances, the outcome of the studies and their interpretation of their relevance to humans is relatively straightforward, indicating that these agents pose or do not pose significant carcinogenic risks to humans. However, in other cases, the interpretation of the results can be quite challenging. In almost all cases, scientists and regulators rely on a weight-of-the-evidence approach, where the number, consistency, and quality of the studies are combined with mechanistic, structure–activity, and other information to reach conclusions about the genotoxicity and likely human carcinogenicity of the agent. In addition to the short-term tests and animal results, information about the genotoxic effects of the pesticide in humans can contribute significantly to the risk assessment process.

11.5  Human biomonitoring To identify pesticides and other agents capable of inducing genotoxicity in humans and to identify groups at elevated risk for cancer or other genetic diseases, biological markers of exposure and effect have been developed to measure genetic changes in exposed humans (Albertini and Hayes, 1997; Albertini et al., 2000; Bolognesi, 2003; Sorsa et al., 1992; Tucker et al., 1997; Wild and Pisani, 1997). These biomarkers range from early premutagenic lesions such as covalent adducts between the chemical and DNA to heritable mutations in endogenous genes such as HPRT. Although these studies have primarily been conducted using somatic cells, a few have been performed using germ cells in which chromosomal changes in human sperm have been monitored. Among the most commonly used biomarkers is the measurement of structural and numerical alterations in lymphocyte chromosomes. In this assay, the frequencies of chromosome changes occurring in metaphase preparations of stimulated peripheral blood lymphocytes from individuals in an exposed group are measured and compared with those of an appropriate control. Increased frequencies of genetic alterations are believed to indicate that an exposure has occurred that is biologically significant and mechanistically related to cancer and other genetic diseases (Sorsa et al., 1992). Consistent with this, studies have shown that individuals with elevated frequencies of structural chromosomal aberrations in their peripheral blood lymphocytes are at increased risk for the development of cancer (Bonassi et al., 1995; Hagmar et al., 1994, 1998). It should be noted that for one frequently measured endpoint, SCEs, such an association was not seen (Hagmar, et al., 1994, 1998). A considerable number of studies have been conducted using various biomarkers to measure genetic alterations in the cells of pesticide-exposed workers. A list of genetic biomarker studies obtained primarily from a search of MEDLINE (and references cited therein) is shown in Table 11.5. Although these studies represent only a fraction of the studies that have been conducted, the results and patterns of response

369

are probably representative of those commonly seen in pesticide biomonitoring studies. As can be seen from the table, numerous reports from many countries have been published on genotoxic effects in pesticide-exposed workers. In many of these, higher frequencies of genotoxic effects have been seen in the exposed workers. However, most studies have been conducted on agricultural workers who have been exposed to many different pesticides. As a result, it is difficult to identify the actual genotoxic agent involved. For example, in the studies conducted in southeast India by Rupa and associates, the cotton field applicators reported having used 11 different pesticides in the period preceding the study (Rupa et al., 1989b). Even in cases in which the exposed workers were exposed primarily to a single pesticide, the reported outcome may be influenced by other confounding factors, such as tobacco smoking, age, exposure to solvents, and inert ingredients. As is also apparent from Table 11.5, studies have been performed for only a small portion of the hundreds to thousands of pesticides currently being used. In addition to the paucity of information on most pesticides, interpreting the results of biomonitoring studies such as these and their significance for workers exposed at lower levels or the general public exposed at much lower levels can be difficult. For example, ethylene oxide has exhibited positive responses in the majority of biomonitoring studies and endpoints measured. This is consistent with the known reactivity of this agent and its results in the short-term genotoxicity tests. This would indicate that at high exposure levels, ethylene oxide poses a genotoxic and carcinogenic risk. However, these studies provide little information about the risk at lower exposure levels, requiring an extrapolation of risk to be made from high exposures to lower exposures. In contrast, negative results were seen in two biomonitoring studies of EDB, an agent that yielded positive results in most short-term tests and was carcinogenic in animals. Although one might interpret these results as indicating that EDB is not genotoxic in humans, this conclusion could easily be in error. In this case, the negative results could simply be due to a combination of low exposures and a limited sample size. Quantitative measures of pesticide exposure are infrequently performed in these types of studies. Other results, such as those reported for DDT, dimethoate, deltamethrin, and cypermethrin, are also challenging to interpret. Based on the short-term test results, one would not expect these agents to be genotoxic. For example, DDT was negative in 133 out of the 143 short-term genotoxicity tests listed in the EPA/IARC GAP database. This suggests that the positive results seen in these types of biomonitoring studies might be due to other factors such as solvent exposure and tobacco use (Petrelli et al., 1993) or may simply be false positives. However, the large number of positive studies in Table 11.5 indicates that pesticide

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370

Table 11.5  Summary of Results of Genotoxicity Studies of Pesticide-Exposed Workers Study group

Location

Pesticide

Endpoint

Result

Reference

Foliage sprayers

Finland

2,4-D and MCPA

SCE

Negative

Linnainmaa (1983)

Workers

United States

DBCP

Sperm aneuploidy

Positive

Kapp et al. (1979)

a

Workers in insecticide plants

Brazil

DDT

Cs aberrations

Positive

Rabello et al. (1975)

Sprayers

Syria

Deltamethrin and cypermethrin

Cs aberrations

Positive

Mohammad et al. (1995)

Accidental exposure of firemen

Brazil

Dimethoate

SCE

Positive

Larripa et al. (1983)

Papaya workers

Hawaii

Ethlylene dibromide SCE

Negative

Steenland et al. (1986)

Papaya workers

Hawaii

Ethlylene dibromide Cs aberrations

Negative

Steenland et al. (1986)

Pesticide sprayers

United States

Ethylene dibromide

SCE

Negative

Steenland et al. (1985)

Pesticide sprayers

United States

Ethylene dibromide

Cs aberrations

Negative

Steenland et al. (1985)

Factory workers

Sweden

Ethylene oxide

SCE

Negative

Hogstedt et al. (1983)

Factory workers

Sweden

Ethylene oxide

Cs aberrations

Positive

Hogstedt et al. (1983)

Factory workers

Sweden

Ethylene oxide

Micronuclei

Negative

Hogstedt et al. (1983)

b

Factory workers

Sweden

Ethylene oxide

Micronuclei

Positive

Hogstedt et al. (1983)

Sanitary workers

Italy

Ethylene oxide

Cs aberrations

Positive

Sarto et al. (1984)

Sanitary workers

Italy

Ethylene oxide

SCE

Positive

Sarto et al. (1984) c

Sterilizer operators

United States

Ethylene oxide

Cs aberrations

Negative

Galloway et al. (1986)

Malathion workers

United States

Malathion

Micronuclei

Negative

Titenko-Holland et al. (1997)

Workers (production)

Czechoslovakia

Mancozeb Cs aberrations containing fungicide Novozir Mn80

Positive

Jablonicka et al. (1989)

Workers (production)

Czechoslovakia

Mancozeb SCE containing fungicide Novozir Mn80

Positive

Jablonicka et al. (1989)

Fumigation workers

Florida

Methyl bromide

Negative

Calvert et al. (1998)

HPRT mutations d

Fumigation workers

Florida

Methyl bromide

Micronuclei

Equivocal

Calvert et al. (1998)

Pesticide plant workers

Brazil

Methyl parathion

Cs aberrations

Negative

de Cassia Stocco et al. (1982)

Pesticide preparing workers

Hungary

Monochlorinated benzene

HPRT mutation

Negative

Major et al. (1992)

Patients (attempted suicide or exposed during work)

Hungary

Organophosphates

Cs aberrationse

Positive

van Bao et al. (1974)

Fumigant applicators

United States

Phosphine

Cs aberrations

Positive

Garry et al. (1989)

Fumigant applicators

United States

Phosphine

SCE

Negative

Garry et al. (1989)

Pesticide applicators

United States

Phosphine

Cs rearrangements

Positive

Garry et al. (1992)

f

Pesticide sprayers

Hungary

Pyrethroids

Cs aberrations

Positive

Nehéz et al. (1988)

Workers (fitters, packers, truck drivers)

Former Soviet Union

Zineb

Cs aberrations

Positive

Pilinskaya (1974)

Store workers and packers

Former Soviet Union

Ziram

Cs aberrations

Positive

Pilinskaya (1970)

Chapter | 11  Genotoxicity of Pesticides

371

Table 11.5  (Continued) Study group

Location

Pesticide

Endpoint

Result

Reference

Pesticide applicators

United States

Herbicides

Cs aberrations

Positive

Yoder et al. (1973)

Pesticide applicators

United States

Insecticides

Cs aberrations

Positive

Yoder et al. (1973)

Farmers

Denmark

Fungicides

Aneuploid sperm

Negative

Harkonen et al. (1999)

Pesticide applicators

United States

Pesticides

Cs aberrations

Positive

Yoder et al. (1973)

Sprayers

New Zealand

Pesticides

SCE

Negative

Crossen and Morgan (1978)

Pesticide workers

Sweden

Pesticides

Cs aberrations

Negative

Hogstedt et al. (1980)

Exposed workers

Hungary

Pesticides

Cs aberrations

Positive

Nehéz et al. (1981)

Agricultural workers

Former Soviet Union

Pesticides

Cs aberrations

Positive

Volnjanskaya (1981)

Floriculturists

Argentina

Pesticides

SCE

Floriculturists

Argentina

Positive

Dulout et al. (1985)

Pesticides

g

Cs aberrations

Negative

Dulout et al. (1985)

f

Greenhouse pesticide sprayers

Hungary

Pesticides

Cs aberrations

Positive

Desi et al. (1986)

Ornamental plant breeders

Argentina

Pesticides

Cs aberrations

Negative

Dulout et al. (1987)

Pesticide workers

Mexico

Pesticides

Cs aberrations

Positive

Gayon et al. (1987)

Pesticides

Cs aberrations

Mixers and field sprayers Hungary

Positive

Paldy et al. (1987)

a

Pesticide sprayers in vineyards

India

Pesticides

Cs aberrations

Positive

Rita et al. (1987)

Pesticide sprayers

Hungary

Pesticides

Cs aberrationsf

Positive

Nehéz et al. (1988)

Vegetable garden workers

India

Pesticides

Cs aberrations

Positive

Rupa et al. (1988)

Fumigant applicators

United States

Pesticides

Cs aberrations

Positive

Garry et al. (1989)

Pesticide sprayers

India

Pesticides

Cs aberrations

Positive

Rupa et al. (1989b)

Pesticide mixers and sprayers

India

Pesticides

Cs aberrations

Positive

Rupa et al. (1989a)

Pesticide applicators

Canada

Pesticides

Micronucleih

Positive

San et al. (1989)

Agricultural workers

Spain

Pesticides

SCE

Negative

Carbonell et al. (1990)

Workers in flower industry

Italy

Pesticides

Cs aberrations

Positive

De Ferrari et al. (1991)

Workers in flower industry

Italy

Pesticides

SCE

Positive

De Ferrari et al. (1991)

Cotton field workers

India

Pesticides

Cs aberrations

Positive

Rupa et al. (1991a)

Pesticide applicators

India

Pesticides

SCE

Positive

Rupa et al. (1991c)

Pesticide applicators

United States

Pesticides

Cs rearrangements

Positive

Garry et al. (1992)

Workers in plastic greenhouses

Greece

Pesticides

Cs aberrations

Positive

Kourakis et al. (1992)

Pesticide workers

Mexico

Pesticides

SCE

Negative

Gomez-Arroyo et al. (1992)

Floriculturists

Italy

Pesticides

Micronuclei

Positive

Bolognesi et al. (1993a)

Floriculturists

Italy

Pesticides

Micronuclei

Positive

Bolognesi et al. (1993b)

Agricultural workers

Spain

Pesticides

Cs aberrations

Positive

Carbonell et al. (1993) (Continued )

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372

Table 11.5  (Continued) Study group

Location

Pesticide

Endpoint

Result

Reference

Agricultural workers

Spain

Pesticides

SCE

Negative

Carbonell et al. (1993)

Pesticide packers

Egypt

Pesticides

Cs aberrations

Positive

Anwar (1994)

Pesticide packers

Egypt

Pesticides

SCE

Negative

Anwar (1994)

Farm workers

Spain

Pesticides

Cs aberrations

Positive

Carbonell et al. (1995)

Greenhouse sprayers

Scandinavia

Pesticides

SCE

Pesticide applicators

India

Pesticides

Negative

Lander and Ronne (1995)

i

Positive

Rupa et al. (1995)

a

Cs aberrations

Dealers and controllers

Syria

Pesticides

Cs aberrations

Positive

Mohammad et al. (1995)

Farmers

Colombia

Pesticides

SCE

Negative

Hoyos et al. (1996)

Farmers

Colombia

Pesticides

Cs aberrations

Negative

Hoyos et al. (1996)

Pesticide sprayers

Greece

Pesticides

SCE

Negative

Kourakis et al. (1996)

Pesticide sprayers

Hungary

Pesticides

Cs aberrations

Positive

Nehéz and Desi (1996)

Farmers

Italy

Pesticides

SCE

Negative

Pasquini et al. (1996)

Farmers

Italy

Pesticides

Micronuclei

Positive

Pasquini et al. (1996)

Greenhouse floriculturists

Italy

Pesticides

DNA adducts

Positive

Peluso et al. (1996)

Greenhouse floriculturists

Italy

Pesticides

Cs aberrations

Negative

Scarpato et al. (1996)

Greenhouse floriculturists

Italy

Pesticides

Micronuclei

Negative

Scarpato et al. (1996)

Greenhouse floriculturists

Italy

Pesticides

SCE

Negative

Scarpato et al. (1996)

Pesticide sprayers

France

Pesticides

DNA damage

Increase

Lebailly et al. (1998)

Mixers and applicators

India

Pesticides

Cs aberrations in spermi

Positive

Rupa et al. (1997)

Pesticide sprayers

Chile

Pesticides

Micronuclei

Negative

Venegas et al. (1998)

Greenhouse workers

Italy

Pesticide

Micronuclei

Positive

Falck et al. (1999)

Pesticide industry workers

India

Pesticides

SCE

Positive

Padmavathi et al. (2000)

Farm workers

Canada

Pesticides

Micronuclei

Equivocal

Davies et al. (1998)

Greenhouse workers

Spain

Pesticides

Micronuclei

Negative

Lucero et al. (2000)

Factory workers

China

Pesticides

Aneuploid sperm

Positive

Padungtod et al. (1999)

Floriculturists

Italy

Pesticides

Micronuclei

Positive

Bolognesi et al. (2002)

Floriculturists

Ecuador

Pesticides

Cs aberrations

Positive

Paz-y-Miño et al. (2002)

Pesticide production workers

India

Pesticides

DNA breakage

Positive

Grover et al. (2003)

Farmers

Portugal

Pesticides

Micronuclei

Positive

Costa et al. (2006)

Farmers

Portugal

Pesticides

SCE

Positive

Costa et al. (2006)

Farmers

Portugal

Pesticides

Cs aberrations

Negative

Costa et al. (2006)

Pesticide production workers

Croatia

Carbofuran

Micronuclei

Positive

Zeljezić et al. (2007)

Chapter | 11

Genotoxicity of Pesticides

373

TABLE 11.5 (Continued) Study group

Location

Pesticide

Endpoint

Result

Reference

Pesticide production workers

Croatia

Carbofuran

DNA breakage

Positive

Zeljezi et al. (2007)

Farmers

India

Pesticides

Cs aberrations

Positive

Naravaneni and Jamil (2007)

Farm workers

Brazil

Pesticides

Micronuclei

Positive

Bortoli et al. (2009)

All studies were performed on peripheral blood lymphocytes unless otherwise noted. a Cs aberrations refer to chromosome and/or chromatid aberrations. b Micronuclei in bone marrow cells. c Negative at low and moderate exposures but positive at high exposures. d Micronuclei in oropharyngeal cells. e Breaks; unstable and stable chromosomal aberrations. f Numerical chromosomal aberrations. g Structural chromosomal aberrations, but exchanges showed a statistically significant increase in exposed over controls. h Micronuclei in exfoliated urothelial cells. i Affecting the 1cen-1q12 region.

exposure is frequently associated with genotoxic effects in exposed workers. Although the vast majority of biomonitoring studies have been conducted using somatic cells, a number of studies have been conducted to measure effects in germ cells (Perry, 2008). Interestingly, positive effects have been reported in many of the studies conducted to date. For example, significant increases in aneuploid sperm were seen in agricultural workers exposed to DBCP (Kapp et al., 1979), in Chinese factory workers exposed to organophosphates (Padungtod et al., 1999), and in Indian applicators and sprayers exposed to a variety of pesticides (predominantly organophosphate insecticides) (Rupa et al., 1997). Moreover, increases in breakage/exchanges affecting the 1cen-1q12 region of chromosome 1 were also detected in the sperm of Indian cotton field workers (Rupa et al., 1997). Notably in earlier studies by Rupa and associates, the Indian group of applicators involved in the sperm and lymphocyte aberration studies had previously been reported to exhibit significant decreases in reproductive performance (fertility, pregnancy loss, and birth anomalies) (Rupa et al., 1991b). These initial reports indicate that exposure to certain pesticides can induce chromosome alterations in the sperm of the exposed workers and may contribute to decreased reproductive performance of the workers.

11.6  Genotoxicity and risk assessment As described previously, short-term tests for genotoxicity are required by regulatory agencies for pesticide registration and play an important role in the safety evaluation and risk assessment process. For the few agents that have been evaluated for heritable risks, genotoxicity assays, particularly those assessing heritable effects in germ cells, have

played a critical role. Historically in cancer risk assessment, the short-term test results and human biomonitoring studies have been used to alert agencies and the public to pesticides with potential cancer-causing properties as well as to provide valuable supplemental information for the positive or negative results seen in animal bioassays. In recently implemented or proposed regulatory strategies, genotoxicity information plays an increasingly important role in the risk assessment process. DNA reactivity and mechanisms of genotoxicity are being used to provide insights into an agent’s mode of action and, as a result, may play a pivotal role in determining whether linear or nonlinear (apparent threshold) models will be used for extrapolation from high animal doses to lower exposure levels. In the U.S. EPA approach, genotoxic effects may also be modeled as precursor events to provide the basis for the selection of a certain extrapolation procedure (Wiltse and Dellarco, 1996). The use of mechanistic or mode of action information plays an important role in the cancer risk assessment guidelines used by the U.S. EPA (U.S. EPA, 2005; Wiltse and Dellarco, 1996) as well as in those implemented by other national and international regulatory groups such as IARC (1999e) and the German Commission for the Investigation of Health Hazards of Chemical Compounds in the Work Area of the Deutsche Forschungsgemeinschaft (Neumann et al., 1998). The evaluation of ethylene oxide provides an example of the contribution of genotoxicity data to the cancer risk assessment process. Upon reviewing the literature on the carcinogenicity of ethylene oxide in humans and animals, the IARC Working Group concluded that there was limited evidence for the carcinogenicity of ethylene oxide in humans but sufficient evidence in animals (IARC, 1994, 2008). However, in its overall evaluation, the IARC Working Group concluded that ethylene oxide is carcinogenic to humans. In making this conclusion, the IARC

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Working Group took into consideration the following evidence (IARC, 2008, p. 287): a. Ethylene oxide is a direct-acting alkylating agent that reacts with DNA. b. Ethylene oxide induces a dose-related increase in the frequency of ethylene oxide-derived hemoglobin adducts in exposed humans and rodents. c. Ethylene oxide induces a dose-related increase in the frequency of ethylene oxide-derived DNA adducts in exposed rodents. d. Ethylene oxide consistently acts as a mutagen and clastogen at all phylogenetic levels. e. Ethylene oxide induces heritable translocations in the germ cells of exposed rodents. f. Ethylene oxide induces a dose-related increase in the frequency of sister chromatid exchange, chromosomal aberrations, and micronucleus formation in the lymphocytes of exposed workers. g. Prospective studies have shown that elevated levels of chromosomal aberrations and micronucleus formation in peripheral blood lymphocytes are associated with increased risks for cancer in humans. In a similar manner with differing conclusions, IARC has evaluated data on the relevance of rodent tumors of the urinary bladder, renal cortex, mammary gland, and thyroid gland induced by agents such as atrazine, chlorothalonil, OPP, p-DCB, and saccharin and their relevance to carcinogenic risk in humans (Rice et al., 1999a). In a number of cases, the lack of genotoxicity exhibited by these agents or their metabolites played an important role in its conclusions that the mechanisms by which agents such as atrazine and saccharin induced cancer in rodents were not relevant to humans (Rice et al., 1999a). Other governmental groups have reached similar conclusions (NTP, 2000; U.S. EPA, 1991, 1998a). It should be emphasized that in all cases, critical evaluation should be used in the interpretation and application of short-term test results in the risk assessment process. Given the large number of tests that can be performed in different cells or strains and at multiple dose levels, positive results should be expected in some tests by random chance alone. As a result, reproducibility and consistency become particularly important in evaluating genotoxicity test results. Short-term tests can also be performed in vitro or in vivo under conditions that will produce positive test results but that are unlikely to pose significant genotoxic risks to humans. For example, there is increasing recognition that positive responses in the in vitro chromosome aberration assay can be caused by mechanisms such as endonuclease activation that are not likely to occur at lower doses (Galloway, 2000; Scott et al., 1991). These tend to occur more frequently at high test concentrations under conditions in which high osmolality, extremes of pH, or excessive cytotoxicity are seen. Similarly, genotoxic effects may

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occur at concentrations in vitro that most likely would not occur in vivo as other types of toxic effects such as neurotoxicity would be dose limiting. A comparison of in vitro concentrations or in vivo animal plasma concentrations with expected plasma levels in humans under conditions of normal (and above normal) usage can assist in the interpretation of the test data. Conversely, negative results in shortterm genotoxicity tests should not be given undue weight because they do not exclude the possibility that an effect occurred in tissues that were not examined, that inadequate bioactivation was used, that the test was improperly conducted, or that the agent induces another type of genetic damage (IARC, 1999e; Proctor et al., 1986). In addition, negative results in these assays cannot be considered to rule out the carcinogenicity of agents that act through other mechanisms (e.g., receptor-mediated effects, cellular toxicity with regenerative proliferation, or peroxisome proliferation) (IARC, 1999e). By using a weight-of-evidence approach to evaluate the data, the likelihood of error (both false positive and false negative) can be minimized. In a similar manner, to confidently use human biomonitoring studies to evaluate risk, one should ensure that the biomarker of interest was sufficiently sensitive to detect changes at the exposure levels of interest, that the number of exposed and control individuals in the study was adequate, that an acceptable number of measurements were collected, and that major confounding variables were controlled. In addition, the identification of the specific pesticide and information on the exposure levels, although frequently difficult to obtain, can add significantly to the evaluation. Although it is uncommon for all of the preceding conditions to be fulfilled, results of human biomonitoring studies, when the specific agent is known, can play a valuable role in the risk assessment process (see the previous example for ethylene oxide).

Conclusion A significant number of pesticides have exhibited genotoxic effects in short-term genotoxicity assays and may pose significant risks to humans. Consistent with this, chromosomal alterations have been seen in many studies monitoring genotoxic effects in pesticide-exposed workers. However, these studies often involve exposures to multiple pesticides and potential confounding factors and have been seen at levels much higher than those experienced by the general public. The ongoing challenge for researchers, regulators, and those interested in environmental health is to effectively use genotoxicity data to distinguish noncarcinogenic and nonmutagenic pesticides from those capable of inducing cancer and heritable mutations in humans, to determine which of the latter pose significant risks at human exposure levels, and, if continued use is needed, to identify safe methods and levels for the use of these agents.

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analysis in peripheral blood lymphocytes and buccal epithelial cells. Mutat. Res. 464, 255–262. Major, J., Kemeny, G., and Tompa, A. (1992). Genotoxic effects of occupational exposure in the peripheral blood lymphocytes of pesticide preparing workers in Hungary. Acta Med. Hungarica 49, 79–90. McCann, J., and Ames, B. N. (1976). Detection of carcinogens as mutagens in the Salmonella/microsome test: assay of 300 chemicals: discussion. Proc. Natl. Acad. Sci. USA 73, 950–954. Mohammad, O., Walid, A. A., and Ghada, K. (1995). Chromosomal aberrations in human lymphocytes from two groups of workers occupationally exposed to pesticides in Syria. Environ. Res. 70, 24–29. Morimoto, K., Sato, M., Fukuoka, M., Hasegawa, R., Takahashi, T., Tsuchiya, T., Tanaka, A., Takahashi, A., and Hayashi, Y. (1989). Correlation between the DNA damage in urinary bladder epithelium and the urinary 2-phenyl-1,4-benzoquinone levels from F344 rats fed sodium o-phenylphenate in the diet. Carcinogenesis 10, 1823–1827. Naravaneni, R., and Jamil, K. (2007). Determination of AChE levels and genotoxic effects in farmers occupationally exposed to pesticides. Hum. Exp. Toxicol. 26, 723–731. National Toxicology Program (NTP) (1986). “Toxicology and Carcino­ genesis Studies of Ortho-Phenylphenol (CAS No. 90-43-7) Alone and with 7,12-Dimethylbenz(a)anthracene (CAS No. 57-97-6) in Swiss CD-1 Mice (Dermal Studies).” NTP, Public Health Service, National Institutes of Health, U.S. Department of Health and Human Services, Research Triangle Park, NC. National Toxicology Program (NTP). (1992). “NTP Technical Report on the Toxicology and Carcinogenesis Studies of Methyl Bromide (CAS No. 74-83-9) in B6C3F1 Mice (Inhalation Studies).” NTP, Public Health Service, National Institutes of Health, U.S. Department of Health and Human Services, Research Triangle Park, NC. National Toxicology Program (NTP) (2000). “Ninth Report on Carcinogens.” NTP, U.S. Department of Health and Human Services, Research Triangle Park, NC. Nehéz, M., and Desi, I. (1996). A genetic toxicological study of pesticide workers. Int. J. Environ. Health Res. 6, 201–208. Nehéz, M., Berencsi, G., Paldy, A., Czeizel, A., Szentesi, I., Levay, K., Maurer, J., and Nagy, E. (1981). Data on the chromosome examinations of workers exposed to pesticides. Regul. Toxicol. Pharmacol. 1, 116–122. Nehéz, M., Boros, P., Ferke, A., Mohos, J., Palotás, M., Vetró, G., Zimányi, M., and Dési, I. (1988). Cytogenetic examination of people working with agrochemicals in the southern region of Hungary. Regul. Toxicol. Pharmacol. 8, 37–44. Neumann, H. G., Thielmann, H. W., Filser, J. G., Gelbke, H. P., Greim, H., Kappus, H., Norpoth, K. H., Reuter, U., Vamvakas, S., Wardenbach, P., and Wichmann, H. E. (1998). Changes in the classification of carcinogenic chemicals in the work area (Section III of the German list of MAK and BAT values). J. Cancer Res. Clin. Oncol. 124, 661–669. Padmavathi, P., Aruna Prabhavathi, P., and Reddy, P. P. (2000). Frequencies of SCEs in peripheral blood lymphocytes of pesticide workers. Bull. Environ. Contam. Toxicol. 64, 155–160. Padungtod, C., Hassold, T. J., Millie, E., Ryan, L. M., Savitz, D. A., Christiani, D. C., and Xu, X. (1999). Sperm aneuploidy among Chinese pesticide factory workers: scoring by the FISH method. Am. J. Ind. Med. 36, 230–238. Paldy, A., Puskas, N., Vincze, K., and Hadhazi, M. (1987). Cytogenetic studies on rural populations exposed to pesticides. Mutat. Res. 187, 127–132.

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Pasquini, R., Scassellati-Sforzolini, G., Angeli, G., Fatigoni, C., Monarca, S., Beneventi, L., DiGiulio, A. M., and Bauleo, F. A. (1996). Cytogenetic biomonitoring of pesticide-exposed farmers in central Italy. J. Environ. Pathol. Toxicol. Oncol. 15, 29–39. Paz-y-Miño, C., Bustamante, G., Sánchez, M. E., and Leone, P. E. (2002). Cytogenetic monitoring in a population occupationally exposed to pesticides in Ecuador. Environ. Health Perspect. 110, 1077–1080. Peluso, M., Merlo, F., Munnia, A., Bolognesi, C., Puntoni, R., and Parodi, S. (1996). (32).P-postlabeling detection of DNA adducts in peripheral white blood cells of greenhouse floriculturists from western Liguria, Italy. Cancer Epidemiol. Biomarkers Prev. 5, 361–369. Perry, M. J. (2008). Effects of environmental and occupational pesticide exposure on human sperm: a systematic review. Hum. Reprod. Update 14, 233–242. Petrelli, G., Siepi, G., Miligi, L., and Vineis, P. (1993). Solvents in pesticides. Scand. J. Work Environ. Health 19, 63–65. Pilinskaya, M. A. (1970). Chromosome aberrations in the persons contacted with ziram. Genetika 6, 157–163. Pilinskaya, M. A. (1974). Results of cytogenetic examinations of people having professional contact with the fungicide zineb. Genetika 10, 140–146. Proctor, B. L., Gaulden, M. E., and Dowd, M. A. (1986). Reactivity and fate of benzene and formaldehyde in culture medium with and without fetal calf serum; relevance to in vitro mutagenicity testing. Mutat. Res. 160, 259–266. Rabello, M. N., Dealmeida, W. F., Pigati, P., Ungaro, M. T., Murata, T., Perira, C. A., and Becak, W. (1975). Cytogenetic study on individuals occupationally exposed to DDT. Mutat. Res. 28, 449–454. Rehner, T. A., Kolbo, J. R., Trump, R., Smith, C., and Reid, D. (2000). Depression among victims of south Mississippi’s methyl parathion disaster. Health Social Work 25, 33. Reitz, R. H., Fox, T. R., Quast, J. F., Hermann, E. A., and Watanabe, P. G. (1983). Molecular mechanisms involved in the toxicity of orthophenylphenol and its sodium salt. Chem.-Biol. Interact. 43, 99–119. Rhomberg, L., Dellarco, V. L., Siegel-Scott, C., Dearfield, K. L., and Jacobson-Kram, D. (1990). Quantitative estimation of the genetic risk associated with the induction of heritable translocations at low-dose exposure: ethylene oxide as an example. Environ. Mol. Mutagen. 16, 104–125. Rice, J. M., Baan, R. A., Blettner, M., Genevois-Charmeau, C., Grosse, Y., McGregor, D. B., Partensky, C., and Wilbourn, J. D. (1999a). Rodent tumors of urinary bladder, renal cortex, and thyroid gland in IARC Monographs evaluations of carcinogenic risk to humans. Toxicol. Sci. 49, 166–171. Rice, J. M., Venitt, S., and McGregor, D. (1999b). The use of short- and medium-term tests for carcinogens and data on genetic effects in carcinogenic hazard evaluation. IARC Sci. Publ. 146, 1–18. Rita, P., Reddy, P. P., and Reddy, S. V. (1987). Monitoring of workers occupationally exposed to pesticides in grape gardens of Andhra Pradesh. Environ. Res. 44, 1–5. Roth, R. (1958). Concerning bronchial cancers in vine-growers injured by arsenic. Virchows Arch. 331, 119–137. Rupa, D. S., Rita, P., Reddy, P. P., and Reddi, O. S. (1988). Screening of chromosomal aberrations and sister chromatid exchanges in peripheral lymphocytes of vegetable garden workers. Hum. Toxicol. 7, 333–336. Rupa, D. S., Reddy, P. P., and Reddi, O. S. (1989a). Analysis of sisterchromatid exchanges, cell kinetics and mitotic index in lymphocytes of smoking pesticide sprayers. Mutat. Res. 223, 253–258.

Chapter | 11  Genotoxicity of Pesticides

Rupa, D. S., Reddy, P. P., and Reddi, O. S. (1989b). Frequencies of chromosomal aberrations in smokers exposed to pesticides in cotton fields. Mutat. Res. 222, 37–41. Rupa, D. S., Reddy, P. P., and Reddi, O. S. (1991a). Clastogenic effect of pesticides in peripheral lymphocytes of cotton-field workers. Mutat. Res. 261, 177–180. Rupa, D. S., Reddy, P. P., and Reddi, O. S. (1991b). Reproductive performance in population exposed to pesticides in cotton fields in India. Environ. Res. 55, 123–128. Rupa, D. S., Reddy, P. P., Sreemannarayana, K., and Reddi, O. S. (1991c). Frequency of sister chromatid exchange in peripheral lymphocytes of male pesticide applicators. Environ. Mol. Mutagen. 18, 136–138. Rupa, D. S., Hasegawa, L., and Eastmond, D. A. (1995). Detection of chromosomal breakage in the 1cen-1q12 region of interphase human lymphocytes using multicolor fluorescence in situ hybridization with tandem DNA probes. Cancer Res. 55, 640–645. Rupa, D. S., Eastmond, D. A., and Reddy, P. P. (1997). Detection of chromosomal alterations in the sperm of pesticide-exposed workers using fluorescence in situ hybridization (FISH). Environ. Mol. Mutagen. 29, 44. San, R. C. H., Rosin, M. P., See, R. H., Dunn, B. P., and Stich, H. F. (1989). Use of urine for monitoring human exposure for genotoxic agents. In “Biological Monitoring for Pesticide Exposure— Measurement, Estimation, and Risk Reduction” (R. G. M. Wang, C. A. Franklin, R. C. Franklin Honeycutt, and J. C. Reinert, eds.), pp. 98–116. American Chemical Society, Washington, DC. Sarto, F., Cominato, I., Pinton, A. M., Brovedani, P. G., Faccioli, C. M., Bianchi, V., and Levis, A. G. (1984). Cytogenetic damage in workers exposed to ethylene oxide. Mutat. Res. 138, 185–195. Sasaki, Y. F., Saga, A., Akasaka, M., Yoshida, K., Nishidate, E., Su, Y. Q., Matsusaka, N., and Tsuda, S. (1997). In vivo genotoxicity of ortho-phenylphenol, biphenyl, and thiabendazole detected in multiple mouse organs by the alkaline single cell gel electrophoresis assay. Mutat. Res. 395, 189–198. Scarpato, R., Migliore, L., Hirvonen, A., Falck, G., and Norppa, H. (1996). Cytogenetic monitoring of occupational exposure to pesticides: characterization of GSTM1, GSTT1, and NAT2 genotypes. Environ. Mol. Mutagen. 27, 263–269. Schmid, R. (1960). Cutaneous porphyria in Turkey. N. Engl. J. Med. 263, 397–398. Scott, D., Galloway, S. M., Marshall, R. R., Ishidate, M. Jr., Brusick, D., Ashby, J., and Myhr, B. C.; International Commission for Protection against Environmental Mutagens and Carcinogens (1991). Genotoxicity under extreme culture conditions. A report from ICPEMC Task Group 9. Mutat. Res. 257, 147–205. Simon, G. S., Tardiff, R. G., and Borzelleca, J. F. (1979). Failure of hexachlorobenzene to induce dominant lethal mutations in the rat. Toxicol. Appl. Pharmacol. 47, 415–419. Slutsky, M., Levin, J. L., and Levy, B. S. (1999). Azoospermia and oligospermia among a large cohort of DBCP applicators in 12 countries. Int. J. Occup. Environ. Health. 5, 116–122. Smith, A. G. (1991). Chlorinated hydrocarbon insecticides. In “Handbook of Pesticide Toxicology” (W. J. J. Hayes and E. R. J. Laws, eds.), pp. 731–915. Academic Press, San Diego. Smith, R. A., Christenson, W. R., Bartels, M. J., Arnold, L. L., St. John, M. K., Cano, M., Garland, E. M., Lake, S. G., Wahle, B. S., McNett, D. A., and Cohen, S. M. (1998). Urinary physiologic and chemical metabolic effects on the urothelial cytotoxicity and potential DNA adducts of o-phenylphenol in male rats. Toxicol. Appl. Pharmacol. 150, 402–413.

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Sorsa, M., Wilbourn, J., and Vainio, H. (1992). Human cytogenetic damage as a predictor of cancer risk. IARC Sci. Publ. 116, 543–554. Steenland, K., Carrano, A., Clapp, D., Ratcliffe, J., Ashworth, L., and Meinhardt, T. (1985). Cytogenetic studies in humans after short-term exposure to ethylene dibromide. J. Occup. Med. 27, 729–732. Steenland, K., Carrano, A., Ratcliffe, J., Clapp, D., Ashworth, L., and Meinhardt, T. (1986). A cytogenetic study of papaya workers exposed to ethylene dibromide. Mutat. Res. 170, 151–160. Sumi, Y., Oode, Y., and Tanaka, H. (2008). Chinese dumpling scare hits Japan – A case of methamidophos food poisoning. J. Toxicol. Sci. 33, 485–486. Suzuki, H., Suzuki, N., Sasaki, M., and Hiraga, K. (1985). Orthophenylphenol mutagenicity in a human cell strain. Mutat. Res. 156, 123–127. Tayama, S., and Nakagawa, Y. (1991). Sulfhydryl compounds inhibit the cyto- and genotoxicity of o-phenylphenol metabolites in CHO-K1 cells. Mutat. Res. 259, 1–12. Tayama-Nawai, S., Yoshida, S., Nakao, T., and Hiraga, K. (1984). Induction of chromosome aberrations and sister-chromatid exchanges in CHO-K1 cells by o-phenylphenol. Mutat. Res. 141, 95–99. Tayama, S., Kamiya, N., and Nakagawa, Y. (1989). Genotoxic effects of o-phenylphenol metabolites in CHO-K1 cells. Mutat. Res. 223, 23–33. Thiers, H., Colomb, D., Moulin, G., and Colin, L. (1967). Cutaneous arsenical cancer in viticultivators in Beaujolais. Ann. Dermatol. Syphiligr. (Paris) 94, 133–158. Titenko-Holland, N., Windham, G., Kolachana, P., Reinisch, F., Parvatham, S., Osorio, A. M., and Smith, M. T. (1997). Genotoxicity of malathion in human lymphocytes assessed using the micronucleus assay in vitro and in vivo: a study of malathion-exposed workers. Mutat. Res. 388, 85–95. Tucker, J. D., Eastmond, D. A., and Littlefield, L. G. (1997). Cytogenetic end-points as biological dosimeters and predictors of risk in epidemiological studies. IARC Sci. Publ. 142, 185–200. U.S. Environmental Protection Agency (EPA) (1979). “Short-Term Tests for Carcinogens, Mutagens and Other Genotoxic Agents,” EPA-625/979-003. U.S. EPA, Research Triangle Park, NC. U.S. Environmental Protection Agency (EPA) (1985). “Health Assessment Document for Chlorinated Benzenes – Final Report,” EPA/600/884/015F. U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (EPA) (1990). “Bromomethane – Integrated Risk Information System (IRIS) Record.” U.S. EPA, Washington, DC. Available at http://www.epa.gov/iris/subst/0015.htm#II. U.S. Environmental Protection Agency (EPA) (1991). “Alpha2u-Globulin: Association with Chemically-Induced Renal Toxicity and Neoplasia in the Male Rat,” EPA/625/3-91/019F. U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (EPA) (1996). “Hexachlorobenzene – Integrated Risk Information System (IRIS) Record.” U.S. EPA, Washington, DC. Available at http://www.epa.gov/iris/subst/0374.htm. U.S. Environmental Protection Agency (EPA) (1997). “1,2Dibromoethane – Integrated Risk Information System (IRIS) Record,” U.S. EPA, Washington, DC. Available at http://www.epa. gov/iris/subst/0361.htm. U.S. Environmental Protection Agency (EPA) (1998a). “Assessment of Thyroid Follicular Cell Tumors,” EPA/630/R-97/002. U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (EPA) (1998b). “Bentazon – Integrated Risk Information System (IRIS) Record.” U.S. EPA, Washington, DC. Available at http://www.epa.gov/iris/subst/0134.htm.

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U.S. Environmental Protection Agency (EPA) (1998c). “Health Effects Test Guidelines OPPTS 870.5100 Bacterial Reverse Mutation Test,” EPA 712-C-98-247. U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (EPA) (1998d). “Health Effects Test Guidelines OPPTS 870.5300 in Vitro Mammalian Cell Gene Mutation Test,” EPA 712-C-98-221. U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (EPA) (1998e). “Health Effects Test Guidelines OPPTS 870.5385 Mammalian Bone Marrow Chromosome Aberration Test,” EPA 712-C-98-225. U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (EPA) (1998f). “Health Effects Test Guidelines OPPTS 870.5395 Mammalian Erythrocyte Micronucleus Test,” EPA 712-C-98-226. U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (EPA) (1998g). “Toxicological Review of Bentazon (CAS No. 25057-89-0): In Support of Summary Information on the Integrated Risk Information System (IRIS).” U.S. EPA, Washington, DC. Available at http://www.epa. gov/iris/toxreviews/0134-tr.pdf. U.S. Environmental Protection Agency (EPA) (1999). “Chlorpyrifos: HED Preliminary Risk Assessment for the Reregistration Eligibility Decision (RED) Document.” U.S. EPA, Washington, DC. Available at http://www.epa.gov/pesticides/op/chlorpyrifos.htm. U.S. Environmental Protection Agency (EPA) (2005). “Guidelines for Carcinogen Risk Assessment,” EPA/630/P-03/001F. U.S. EPA, Washington, DC. Ushiyama, K., Nagai, F., Nakagawa, A., and Kano, I. (1992). DNA adduct formation by o-phenylphenol metabolite in vivo and in vitro. Carcinogenesis 13, 1469–1473. van Bao, T., Szabo, I., Ruzicska, P., and Czeizel, A. (1974). Chromosome aberrations in patients suffering acute organic phosphate insecticide intoxication. Humangenetik 24, 33–57. Venegas, W., Zapata, I., and Marcos, R. (1998). Micronuclei analysis in lymphocytes of pesticide sprayers from Concepcion, Chile. Teratogen. Carcinogen. Mutagen. 18, 123–129.

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Volnjanskaya, V. (1981). Level of chromosome aberrations in agricultural workers. Gigiena Truda i Professional ‘nye Zabolevaniia 12, 47–48. Waters, M. D., Stack, H. F., Garrett, N. E., and Jackson, M. A. (1991). The Genetic Activity Profile database. Environ. Health Perspect. 96, 41–45. Waters, M. D., Stack, H. F., and Jackson, M. A. (1999). Short-term tests for defining mutagenic carcinogens. IARC Sci. Publ. 146, 499–536. Wei, L. Y., Chao, J. S., and Hong, C. C.. (1997). Assessment of the ability of propoxur, methomyl, and aldicarb, three carbamate insecticides, to induce micronuclei in vitro in cultured Chinese hamster ovary cells and in vivo in BALB/c mice. Environ. Mol. Mutagen. 29, 386–393. Whorton, D., Milby, T. H., Krauss, R. M., and Stubbs, H. A.. (1979). Testicular function in DBCP exposed pesticide workers. J. Occup. Med. 21, 161–166. Wild, C. P., and Pisani, P. (1997). Carcinogen-DNA and carcinogen-protein adducts in molecular epidemiology. IARC Sci. Publ. 142, 143–158. Wiltse, J., and Dellarco, V. L. (1996). U.S. Environmental Protection Agency guidelines for carcinogen risk assessment: past and future. Mutat. Res. 365, 3–15. Yoder, J., Watson, M., and Benson, W. W. (1973). Lymphocyte chromosome analysis of agricultural workers during extensive occupational exposure to pesticides. Mutat. Res. 21, 335–340. Zahm, S. H., and Ward, M. H. (1998). Pesticides and childhood cancer. Environ. Health Perspect. 106(suppl 3), 893–908. Zahm, S. H., Ward, M. H., and Blair, A. (1997). Pesticides and cancer. Occup. Med. 12, 269–289. Zeljezi, D., Vrdoljak, A. L., Radić, B., Fuchs, N., Berend, S., Orescanin, V., and Kopjar, N. (2007). Comparative evaluation of acetylcholinesterase status and genome damage in blood cells of industrial workers exposed to carbofuran. Food Chem. Toxicol. 45, 2488–2498.

Chapter 12

Developmental and Reproductive Toxicology of Pesticides Poorni Iyer1,* and Susan Makris2 1

California Environmental Protection Agency, Office of Environmental Health Hazard Asessment, Sacramento, California United States Environmental Protection Agency, National Center for Environmental Assessment, Office of Research and Development, Washington, D.C. 2

12.1  Introduction Pesticides are chemicals deliberately used in households and modern agriculture, and subject to much regulation. As a class they typically are well studied especially due to growing concern about the safety of pesticides and how exposure may affect human health. Exposure either occupational or via food residues/contamination of air and water and reproductive outcome in populations has spurred a lot of the attention. Several studies in children and pregnant women using urine, amnionic fluid and meconium have demonstrated that from 89% to 100% of fetuses in the United States are exposed to pesticide in utero and most are exposed to mixtures of several pesticides (Whyatt and Barr, 2001). Additionally, the National Health and Nutrition Examination Survey (NHANES) found that 95% of the U.S. population has measurable pesticide metabolites in urine samples (Barr et al., 2004). Increasing demand for organically grown produce (as defined by The Organic Foods Act of California, 1990), and reports on the levels of pesticides in the diets of infants and children (NRC, 1993) and the passage of federal regulations in the United States, e.g. the Food Quality Protection Act (U.S. 104th Congress, 1996), has shifted the focus to the efficacy of the chemicals used while weighing the risk to human health and environmental impact. Human malformations occur in roughly 5% of live births, therefore to demonstrate an increase in the overall rate of malformation or incidence of a specific type of malformation from a documented exposure, a much larger population *

The views expressed in this chapter are the authors’, and they do not necessarily reflect the policies of the California EPA or U.S. EPA.

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is required than if the background rate were zero (Fraser, 1977). The limitations often associated with epidemiological data, such as recall bias, lack of specificity and use of surrogates for exposure, must be considered in evaluating findings that suggest an association with pesticide exposure. One explanation for this is the specificity of the compounds involved in the exposure. Given that exposure is often categorized as either general pesticide handling or agricultural setting, exposure to specific compounds is not evaluated. Adverse effects of pharmaceutical agents have been predicted from data on laboratory animals at exposures near maternally toxic levels (Johnson et al., 1990). Much of the animal data on the reproductive and developmental effects of agrochemicals are generated for the purpose of pesticide registration to meet the regulations of the Federal Insecticide Fungicide and Rodenticide Act, FIFRA (U.S. EPA, 1982), and may not be very accessible in the open literature. Since the amount of published information is limited, an attempt is made to fill that void by addressing several aspects in the area of developmental and reproductive toxicity of agrochemicals focusing primarily on pesticides. Typically studies on pesticide use and pregnancy outcome generally concentrate on birth defects and the effects on the reproductive system in most agriculture-related occupations. Early and late fetal loss, alteration in gestational age at delivery, formation of terata (birth defects), infant/child morbidity and mortality, male/female sexual dysfunction, sperm abnormalities, amenorrhea, dysmenorrhea, illness during pregnancy and parturition, and endocrine effects are included in the endpoints examined. A number of epidemiological studies have some indications of elevated risk, and the degree to which pesticide exposure may or may not be responsible for developmental problems in humans is receiving increasing 381

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attention. The Pesticide Residue Monitoring Program in Europe has provided data for indirect exposure asseement to endocrine disruptors in the general population (Mantovani et al., 2008). Such indirect exposure assessments through food and environmental programs help in monitoring exposure trends and identifying priorities for biomonitoring. While accurate health risk assessments may not be determined, biomonitoring studies can provide substantial opportunities with known exposure levels for human samples. Pesticides acting on the developing organism or on the reproductive system may produce adverse effects by one of several mechanisms. They may be direct-acting by being chemically reactive and (1) cause germ cell destruction (e.g. alkylating agents), or (2) exert their effects due to their structural similarity to endogenous molecules, e.g. hormone agonists/antagonists such as phytoestrogens. They could also act indirectly and interrupt reproduction (1) by metabolism to a direct-acting compound or reactive intermediate, (2) via endocrine alterations such as increased/decreased steroid clearance, or (3) by stimulating or inhibiting neuroendocrine responses at the level of the thyroid, hypothalamus or pituitary. Developmental toxicants, through a direct- or indirect-acting mechanism, may result in either embryolethality, frank malformations or other undesirable sequelae such as growth retardation or functional alteration. Similarly, pesticides affecting reproduction may act on selected stages targeting the prenatal stage, prepubertal stage or the adult, resulting in damage to the reproductive organs and/or impaired fertility. The potential of pesticides to adversely affect development is determined from epidemiological data or from studies conducted in laboratory animals to meet the regulations of the Federal Insecticide Fungicide and Rodenticide Act, FIFRA (U.S. EPA, 1982), per testing guidelines published by regulatory agencies and authorities such as the U.S. EPA and the Organization for Economic Co-operation and Development (OECD) (OECD, 1983, 2001a,b). Included in this chapter is information from experimental studies on developmental toxicity and reproductive toxicity for the major classes of pesticides. A brief account of some of the studies (published or from regulatory agency reviews that may be accessed via the U.S. Freedom of Information Act [FOIA]) including species, dose, gestational age at the time of exposure and the type of effects observed are presented. The references were selected based on their originality, availability of an adequate review, or because they were current at the time of preparing the manuscript.

12.1.1  Developmental Toxicity Studies covering the period from conception to the completion of morphological structure and functional capability of the individual are included under this category. Malformations caused by the drug thalidomide in the early 1960s have largely been instrumental in investigating the

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role of chemical exposure in the causation of birth defects. At that time, the U.S. Food and Drug Administration (FDA) established requirements for preclinical testing of pharmaceuticals, and the protocols developed served as the template for the conduct of animal studies used to examine the effects of pesticides. Findings from such studies are extrapolated to humans and include malformations in fetuses, as well as end points such as prenatal death, growth alterations, developmental variations, and effects on postnatal development. Adverse outcomes that result from preconceptional or developmental exposures may be observed immediately, or they may be expressed as latent effects that are not evident until later in life (Selevan et al., 2000; WHO, 2007). Since a number of organ systems continue to develop postnatally, consideration of both structural and functional development can be important for an adequate assessment of developmental toxicity. Since the passage of regulations intended to ensure adequate safety assessment of pharmaceuticals intended for pediatric use (i.e. the FDA Modernization Act Pediatric Exclusivity Provision, 1997; the Pediatric Rule, 1998; the Best Pharmaceuticals for Children Act, amending FFDCA Section 505A, 2002; and the Pediatric Research Equity Act, 2003), the U.S. FDA has focused on the conduct and assessment of juvenile animal toxicity studies in preclinical developmental toxicity testing paradigms. The postnatal testing paradigm must be designed on a case by case basis, considering such aspects as the intended use of the pharmaceutical, any known target organ toxicity, and the developmental stage of concern (U.S. FDA, 2006). In concept, the U.S. EPA has embraced this approach to assessment of postnatal developmental outcomes (Hurtt et al., 2004). Currently, the only standardized study protocol that addresses postnatal functional outcomes for pesticide hazard assessment is the developmental neurotoxicity study. This study is considered particularly important for pesticides, since alterations in postnatal neurodevelopment and behavioral changes have been reported in laboratory animals for a number of pesticides (e.g. Boyes et al., 1997; Chernoff et al., 1979a; Gray et al., 1986; Ostby et al., 1985; Sette, 1989; Tilson, 1998). Increased perinatal mortality has been reported in laboratory animals from excessive exposure to pesticides; in one case the deaths may have been related to functional cardiac disorders (Grabowski and Daston, 1983). In addition, postnatal exposure via lactation has resulted in the induction of cataracts in rat pups (Chernoff et al., 1979b; Gaines and Kimbrough, 1970). Epidemiological studies have documented an association of spontaneous abortions and fetal deaths with maternal exposure (Barlow and Sullivan, 1982; Goulet and Theriault, 1991; Weinberg, 1993; Wilson, 1979). The potential of pesticides to cause adverse effects on the developing individual has been demonstrated in laboratory animal studies with a range of effects at various stages of development. Of the numerous pesticides tested so far, about 43% have been documented to induce birth defects in experimental animals (Schardein,

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1993). All chemicals can interfere with some aspect of development if administered at a sufficiently high dose level at the appropriate time of development to certain species of animal, and hence the findings from regulatory studies provide information on the potential of the chemical to result in adverse effects on the conceptus. Development is a complex process and the effects of a chemical depend on the time of exposure, the exposure level, and the extent of maternal effects. Often the nature of an insult during embryonic development is less important than the developmental stage at which it occurs. This is because the steps in the sequence of tissue interactions during development are susceptible to disruption for a specific period of time. Often, early exposure, i.e. during preimplantation and early postimplantation, results in fetal death, while exposure during the organogenesis period (3 weeks after conception through 2 months in humans) results in structural birth defects. However, there is evidence that exposure during the preimplantion period can also result in teratogenic effects (Rutledge et al., 1992; Spielmann and Vogel, 1989). Pre- or postimplantation exposure of the developing conceptus to toxicants may also result in a “derailment” in the genetic control of development and the coordinated cascade of events that occur during normal development. Thus developmental abnormalities may be induced by disrupting the coordinated expression of developmental genes involved in genomic imprinting, cell lineage specification, cell mixing and recognition, cell–cell interaction, cell migration and differentiation, and segmentation, depending on the time of exposure (Kimmel et al., 1993). Exposure after the critical stage of organogenesis often results in growth retardation or other functional deficits. For regulatory purposes, hazard identification is based on the dose level at which an effect is noted, the observation of a dose response and, also, on whether the adverse effect on the conceptus occurs at an exposure level below that which causes severe maternal toxicity. This is done partly to determine if the maternal effects are the underlying mechanism for the developmental effects noted. Testing for developmental toxicity therefore requires the use of relatively high doses even though humans may actually be exposed to lower environmental concentrations of the pesticide. The details on testing protocols will be elaborated later in this chapter, suffice it to say at this point that the purpose for testing at high doses is to gain an understanding of the mechanism of action of the chemical or to detect rare events in a study with limited statistical power. The limitations of such testing, as in other toxicology studies, include the range of sensitivity within humans, extrapolation of effects observed at high doses to predict those likely to occur at low doses, and extrapolation from tests in animals to humans.

12.1.2  Reproductive Toxicity Adverse effects that constitute reproductive toxicity range from small decrements in reproductive ability in either

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males or females to a state of functional infertility. It also includes effects on the reproductive organs irrespective of the influence on fertility in the affected individual. This is particularly relevant when animals are used as models for effects in humans because fertility in rodents is often difficult to disrupt and other indicators of reproductive function may be more sensitive. Hence, fertility cannot be used as the only tool to diagnose adverse effects. The effects of pesticides on reproduction may be acute or chronic and may be directed to a single sex. Also, pesticides affecting reproduction may act on selected stages targeting the prenatal stage, prepubertal stage, or the adult. A few examples of reproductive toxicity by one or more of these mechanisms will be reviewed herein. Furthermore, due to unique anatomical and physiological characteristics in different species, the effects noted may differ. A pesticide causing reproductive toxic effects in one species may not be toxic in another and hence the relevance to humans needs to be examined. The impact on fertility also needs to be considered, since the level of sperm production differs across species and a decrease in sperm count may not have the same impact in all species. Effects involving the endocrine system may affect the developing reproductive system, and while compounds implicated in such phenomena are often considered endocrine receptors, they are essentially reproductive toxicants. Recent regulatory efforts under FQPA have targeted such chemical effects in ensuring the safety of exposed populations.

12.1.3  Epidemiology The incidence of pesticide-related adverse reproductive/ developmental outcomes has been extensively reviewed (Sever 1988; Sever et al., 1997; Weselak et al., 2007). A number of epidemiological studies have linked pre- and postnatal exposures to pesticides to a number of adverse developmental outcomes, including fetal death, intrauterine growth restriction, preterm birth, and birth defects (Weselak et al., 2007). The degree to which pesticide exposure may or may not be responsible for developmental problems in humans is often unclear, because many of the studies rely on job title only and/or the exposure category “any pesticide” as a measure of exposure. Thus the evidence is often limited or inadequate evidence to support causality for all associations examined. Research on organophosphate pesticide exposure and neurodevelopment suggests some negative association of exposure and neurodevelopment at certain ages and increased levels of organophosphate exposure in utero result in greater numbers of abnormal reflexes in neonates and studies in older infants and young children. In young children (2–3 years) two separate studies observed an increase in maternally reported pervasive developmental disorder with increased levels of organophosphate exposure (Rosas and Eskenazi, 2008). Elevated risk of limb anomalies (Lin, 1994; Schwartz and LoGerfo, 1988; Schwartz

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et al., 1986) has been associated with ecological exposure and occupational exposure; and orofacial clefts (Nurminen, 1995) have been related to maternal environmental exposure. In several countries it has been noted that maternal agricultural occupation and pesticide exposure may be associated with elevated risk of spontaneous abortion and stillbirth (Goulet and Theriault, 1991; Heidam, 1984; Restrepo et al., 1990a, b; Rita et al., 1987). Epidemiological studies have also linked pre- and postnatal exposures to pesticides with neurological deficits (Garry et al., 2002a; Guillette et al., 1998). While reports from accidental exposure as well as occupational use have documented that pesticides can be incriminated in adverse reproductive outcomes, some studies have found no indication of reproductive hazards, presenting rather inconclusive results (Nurminen, 1995). Exposure to organochlorine and organophosphate pesticides in grape gardens of India resulted in higher abortion rates (almost sixfold) in 12 exposed couples compared to 15 nonexposed couples (Rita et al., 1987). The compounds handled in this study included DDT, lindane, quinalphos, dithane M45, metasystox, parathion, copper sulfate, dichlorovos, and dieldrin. Similarly, women working in vineyards in Crimea also had higher rates of miscarriage after exposure to DDT, sulfur, methyl parathion and copper sulfate (Nikitina, 1974). In China, women exposed to chlorophenamidine (chlordimeform), dikishuang, and kitazin were found to be at increased risk of delivering stillbirths (RR  1.4–1.81) as well as spontaneous abortions (RR  1.90–4.00 depending on gravidity). The risks would be even higher, since previous adverse pregnancy outcomes were controlled for in the study (Weinberg, 1993). In rural California, second trimester occupational exposure to pesticides was associated with an odds ratio of 4.8 in a case-control study of stillbirths and early neonatal deaths (Pastore et al., 1995). Male and female farmers exposed to pesticides in central Sudan had higher odds ratio for stillbirth (500 g) in a case-control study; OR  5.1, 95% CI  1.4–9.6 (Taha and Gray, 1993). Fetal death (Arbuckle and Sever, 1998), intrauterine growth restriction, preterm birth (Longnecker et al., 2001; Savitz et al., 1997), and birth defects (Correa-Villasenor et al., 1991; Rupa et al., 1991; Sever et al., 1997) have also been linked to pesticide exposure.

12.2  Exposure As mentioned previously, human malformations occur in roughly 5% of live births. Therefore, to demonstrate an increase in the overall rate of malformations or the incidence of a specific type of malformation from a documented exposure, a much larger population is required than if the background rate were zero (Fraser, 1977). In evaluating findings that suggest an association with pesticide exposure, limitations that are often associated with epidemiological data, such as recall bias, lack of specificity, and use of surrogates for exposure, must be considered. While the specificity of the

Hayes’ Handbook of Pesticide Toxicology

compounds involved in the exposure is typically not available, the effects of known classes of chemicals can be studied, since a number of compounds have a common mechanism of action as reviewed in this chapter. The importance of interactions between genetic susceptibility and in utero pesticide exposure has also been reported (Berkowitz et al., 2004). Data on laboratory animals at exposures near maternally toxic levels (Johnson et al., 1990) have been used to predict adverse effects of pharmaceutical agents and this approach has been used in evaluating the safety of drugs to women of child-bearing age. Similarly, data on the reproductive and developmental effects of pesticides are generated in laboratory animals for the purpose of pesticide registration under FIFRA. These data typically are unavailable in the open literature and the amount of published information is limited. Unlike review articles, this chapter will include findings from such studies on developmental and reproductive toxicity of pesticides.

12.2.1  Timing of Exposure Timing influences the various developmental effects that are observed since peturbations at specific windows of development can have adverse impacts on the process underway (Selevan et al., 2000). The concept that insult prior to the beginning of “organogenesis” results only in an “all (i.e. death) or none” effect is no longer considered accurate. Abnormal development subsequent to insult at preimplantation stages suggestive of early alterations in pattern formation has been reported for retinoic acid (Rutledge et al., 1994). More regarding developmental stage-related sensitivities and this area of pattern formation and early alterations is of concern in the area of pesticide exposure. Just as the time of exposure determines the developmental effects of a chemical, toxicity to the reproductive system also varies with the timing of exposure. Accordingly, reproductive toxicants have been classified as described below.

12.2.2  Prenatal Reproductive Toxicants These are compounds that affect the developing reproductive system in utero resulting in prenatal ovarian or testicular toxicity in humans and animals. These include absence of or a considerable decrease in the number of primordial oocytes (e.g. primary or secondary amenorrhea). Thus while it is possible that prenatal exposure could affect the oocyte, current study protocols are not designed to detect subtle changes that may occur. More frequent testing for toxicity to male reproductive processes is conducted because of the premise of male sensitivity and the ease of access to gametes and gonads. Also it is often presumed that the female gamete is better protected from mutagenic chemicals due to the probability that chemically induced DNA damage in a primary oocyte is repaired prior to ovulation (Preston et al., 1995).

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

Despite the differences in males and females in anatomy and biological control mechanisms for reproduction, in the absence of data to the contrary, it should be assumed that both male and female gametes are equally sensitive to reproductive toxicants. This is particularly true given the recent evidence that mammals may produce egg cells after birth (Zou et al., 2009).

12.2.3  Prepubertal Reproductive Toxicants The modulation of the hypothalamic–pituitary–ovarian axis is influenced by higher centers in the central nervous system. Both ovulation and ovarian hormone production require the interaction of components of this axis and the effects of specific compounds on any of these levels exert their effect on reproduction. The effects can be elucidated clinically by examining the impact on menstrual cyclicity in humans and estrus cyclicity in nonprimate animals. While there are no data to link the increasing trend of early menarche with pesticide use, the area of estrogenic effects and their role on the onset of puberty is receiving attention (Thigpen et al., 1999). In the rat, studies on dams consuming diets containing high concentrations of estrogenic substances, with resultant exposure of their pups in utero and prior to weaning, suggest that the estrogen imprinting metabolism of the pups or future responses to other exogenous estrogenic substances may be altered (Lamartiniere et al., 1995). Thus the effects of exposure may be noted in subsequent generations since a number of pesticides may have estrogenic potential. The prepubertal gonad may differ from the sexually mature gonad in its sensitivity to the toxic impact of pesticides and this is an endpoint that deserves examination. Contaminants of pesticides, such as TCDD, may in fact have such an effect, but the findings are not conclusive. There is increasing attention being paid to the latent effects of pesticides on sexual differentiation in rodents.

12.2.4  Adult Reproductive Toxicants The effects on the reproductive system may be observed in the adults as well as in their progeny if exposure occurs over a long period of time. These are generally detected in the multigeneration studies conducted in laboratory animals. Studies submitted for regulatory purposes (e.g. fenthion, oxydemeton methyl) have demonstrated effects such as increased epididymal vacuolation, and other histopathological changes (CDPR [California Department of Pesticide Regulation] DATABASE, 2000). Gender differences in response to chemical insult must be taken into consideration, e.g. while damage to spermatocytes in the male may have transient adverse effects on reproductive capacity, damage to oocytes is permanent. The reproduction study guidelines (OECD, 2008; U.S. EPA, 1998a,b,c) require both qualitative and quantitative ovarian histo­ pathology data to detect changes that may be occurring due

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to prolonged pesticide exposure over various developmental periods. Additionally, recent studies have documented adult/pubertal alterations resulting from gestational and/or neonatal exposures (Gray and Kelce, 1996). Hence studies should include a comprehensive assessment of reproductive function after perinatal exposure because the developing animal may be extremely sensitive to toxicants during sex differentiation, and a number of these effects are difficult to detect until late in life.

12.3  Mechanisms of action Compounds used as pesticides have different mechanisms of action and these may be independent of the species targeted. Pesticides can therefore be studied by their mechanism of action.

12.3.1  Direct-Acting Pesticides that are direct-acting may exert their effect by being chemically reactive; these compounds may be nonspecific in their site of action and most of these compounds are cytotoxic, carcinogenic, mutagenic, or developmentally toxic. They may also be toxic to the reproductive system, and, in fact, the disruption of reproductive function could occur at doses lower than those which cause tumors. The classic example of such a mechanism is the case where the risk of sterility following many forms of cancer chemotherapy is considerably higher than the risk of second tumors (Kay and Mattison, 1985). Other direct-acting compounds are structurally similar to endogenous molecules, such as some organochlorines which may exert their effects through interaction with estrogen receptors. Organochlorines have been implicated in abnormal menses and impaired fertility (Mattison et al., 1983).

12.3.2  Indirect-Acting Developmental/reproductive toxicants that are metabolized to either chemically reactive products or structures similar to endogenous molecules fall into this group. The embryo, fetus as well as both the ovary and the testis have been demonstrated to have microsomal monooxygenases, epoxide hydrases, and transferases responsible for metabolizing xenobiotics (Dixon and Lee, 1980; Heinrichs and Juchau, 1980; Mattison and Thorgeisson, 1978, 1979; Pedersen et al., 1985).

12.4  Regulatory issues Outlined in this section are the many issues dealing with the use of pesticides and their regulation by state, national and international agencies. These include the conduct and interpretation of studies as well as the application of new findings and regulations.

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12.4.1  History The effects of thalidomide and the Kefauver-Harris Act in 1962 led the Food and Drug Administration (FDA) of the U.S. government to strengthen drug testing. Currently, the U.S. government requires manufacturers to perform hazard assessments to determine the teratogenic potential of chemicals. The U.S. EPA published teratogenicity testing requirements in 1978 under FIFRA. Essentially it mandated how testing was to be conducted and reported, which differed little from the FDA guidelines, except that exposure was to be initiated just before implantation and concluded the day before delivery (U.S. EPA, 1978).1 In the early 1980s, the U.S. EPA specified the kinds of data and information required under FIFRA to support the registration of pesticides (U.S. EPA, 1982), reflecting guidelines proposed in 1978 (U.S. EPA, 1978, 1984). These testing requirements, for both food-use and nonfood-use pesticides were revised and updated in 2007 to reflect advances in science and risk assessment that had occurred over the intervening years (U.S. EPA, 2007). Similar regulations also went into effect through the EPA for chemicals under the Toxics Substances Control Act (TSCA); these were revised in 1985 (U.S. EPA, 1985). In 1986 the EPA published procedures to evaluate potential developmental toxicity associated with human exposure to environmental toxicants (U.S. EPA, 1986). Also, a screening test for developmental neurotoxicity to include behavioral and neuropathology analyses was proposed (Francis, 1987; U.S. EPA, 1986) and finalized into test rules in 1988 and 1989 (U.S. EPA, 1988, 1989). Postnatal functional assessment has been recognized as an important part of developmental toxicity testing in the United States and is required in some cases. In other countries, requirements are in place for behavioral testing as a part of developmental toxicity testing (Barlow, 1985; EEC, 1983; Tanimura, 1985; WHO, 1986). In November 1986, voters in the state of California approved an initiative to address concerns about exposure to toxic chemicals. That initiative became the Safe Drinking Water and Toxic Enforcement Act of 1986, better known as Proposition 65. This requires the Governor to publish a list of chemicals that are known to the state to cause cancer, birth defects, or other reproductive harm. The chemicals that cause birth defects or other reproductive harm are called reproductive toxicants. The Proposition 65 list contains a wide range of chemicals, including dyes, solvents, pesticides, drugs, and food additives. If a pesticide is on the list the employer must warn the employee if the exposure levels of the pesticide present a significant health risk; the employer may also choose to provide a warning simply based on the presence of the chemical, even if the risk is not significant. In the 1

While the FDA requires three studies covering different segments of development, the EPA requires (1) a standard prenatal developmental toxicity study with exposure during the main period of organogenesis and (2) a two-generation reproduction study.

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case of worker exposure to pesticides, this warning is provided through the required hazard communication procedures, and as an agricultural crop producer, the employer is also required to keep application-specific information on the pesticides used. A number of pesticides have been listed and subsequently withdrawn from registration for use in the state. A complete list of compounds may be accessed via the internet at url: http://www.oehha.ca.gov/prop65/prop65_list/files/ P65single080709.pdf/.

12.4.2  Principles of Testing and Evaluation In 1998, the EPA Office of Prevention, Pesticides and Toxic Substances (OPPTS) revised the 1982 Health Effects Test Guidelines, (U.S. EPA, 1998), simultaneously harmonizing with guidelines of the Organization for Economic Co-operation and Development (OECD) which were also undergoing revision and update. The OPPTS harmonized guidelines have been developed for use in the screening and testing of pesticides and toxic substances; they describe protocols for the collection of laboratory animal test data that must be submitted to the Agency for review under Federal regulations. The purpose of harmonizing these diverse guidelines into a single unified set of OPPTS guidelines was to minimize variations among the testing procedures that must be performed to meet the data requirements of both the FIFRA (7 U.S.C. 136, et seq), as amended by the Food Quality Protection Act (FQPA) (Pub. L.104–170), and the Toxic Substances Control Act (TSCA) (15 U.S.C. 2601). The OPPT guidelines under 40 CFR 798.4900 and 40 CFR 798.4700, OPP guidelines 83-3 and 83-4, and OECD guidelines 414 and 416 provided the source material for developing these harmonized OPPTS test guidelines. Revisions to the previous set of established testing guidelines (U.S. EPA, 1984) were considered over a period of years, and the process incorporated the involvement of industry, multiple U.S. and international government agencies and organizations, non-governmental organizations (NGOs), public advocacy groups, public comment, and U.S. EPA Science Advisory Board and FIFRA Scientific Advisory Panel reviews. Some of the important improvements in the prenatal developmental toxicity study design include an increase in the number of animals (litters) assigned to testing in the developmental toxicity study conducted in rabbits, and an extension of the exposure period to include maternal dosing from the time of conception or implantation until the day prior to pregnancy termination. In the multigeneration reproduction study design, a number of important endpoints were added, including sperm/semen evaluation, examination of vaginal smears for evaluation of estrous cyclicity, expanded organ weight and histopathology data of reproductive, endocrine, and other target organs, measures of anogenital distance (triggered by other adverse findings, e.g. alterations in offspring sex ratio), and an evaluation of the age of sexual maturation in second generation offspring (balanopreputial separation in males and vaginal

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

patency in females) (Kimmel and Makris, 2001). Many of these improvements to the reproduction study guideline are valuable in the assessment of effects resulting from endocrine disruption. The standardized evaluation of the effects of endocrine disrupting chemicals on development and reproduction, which will be implemented by the U.S. EPA in accordance with the FQPA amendments to Section 408(p) of the Federal Food, Drug, and Cosmetic Act (FFDCA), has been developed through a number of national and international collaborative efforts. Protocol validation data, a proposed tiered testing paradigm, and a recommended list of chemicals to test utilizing the validated protocols have been reviewed by a U.S. EPA Science Advisory Board and FIFRA Scientific Advisory Panel (U.S. EPA, 1999, 2009a,b). The testing paradigm focuses on the estrogen, androgen, and thyroid endocrine systems, and utilizes a variety of in vitro, ex vivo, and in vivo assays. For those chemicals that raise positive signals regarding their endocrine-active potential, a full two-generation reproduction study is required as a second tier test. Overall, the guidelines provide information on the appropriate study design and methodology for the conduct of studies and may be publicly accessed via the U.S. EPA and OECD websites (see the individual guideline citations in the References). Tables 12.1 and 12.2 describe the protocols for the developmental and reproductive toxicity testing and note the approximate dosing and breeding schedules. Other testing paradigms are also utilized for screening and prioritization of pesticide inerts as well as high production volume (HPV) chemicals by the U.S. EPA and OECD. These programs include the conduct of studies that evaluate some aspects of development and reproduction. Harmonized guidelines exist for the recommended studies, i.e. the Combined Repeated Dose Toxicity Study with the Reproduction/Developmental Toxicity Screening Test (OPPTS 870.3550; OECD GL 422) (OECD, 1996; U.S. EPA, 2000b) and the Reproduction/Developmental Toxicity Screening Test (OPPTS 870.3650; OECD GL 421) (OECD, 1995; U.S. EPA 2000a,b). The value and use of these tests is somewhat limited, and they are not considered to be adequate for a full characterization of developmental or reproductive toxicity following pesticide exposure (OECD, 2008). For a number of organ systems, the critical period for inducing abnormalities may extend to the postnatal period, e.g. renal development (Couture, 1990). To some extent, it is expected that some functional deficiencies and other postnatal effects will be detected in reproduction (two-generation) studies. However, that approach is not practicable given the many organ systems that might be assessed, each with their own inate developmental timing, and structural and functional endpoints that might be perturbed by chemical exposure. As previously mentioned, the only standardized study protocol that addresses postnatal functional outcomes for pesticide hazard assessment is the developmental neurotoxicity study,

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Table 12.1  Protocol fo elopmental Toxicity Testing Rats: 20 females per dose group Day 0 (evidence of sperm/plug in female or in bedding)

Begin exposure (around implantation) generally on day 0 or 6 and continue until day before parturition

Day 21

C-section and examine fetuses (generally C-section is on gestation day 20 if mating is on gestation day 0)

Rabbits: 20 females   per dose group Day 0 (day of artificial insemination; natural mating can be used)

Begin exposure (around implantation) generally gestation day 0 or 7 and continue until day before parturition

Day 29

C-section and examine fetuses

Notes: At minimum, the test substance should be administered daily from around the time of implantation to the day before caesarean section on the day prior to the expected day of parturition. Alternatively, if preliminary studies do not indicate a high potential for preimplantation loss, treatment may be extended to include the entire period of gestation, from fertilization to approximately 1 day prior to the expected day of parturition. It is preferred that the dams are exposed from the time of mating. The timing of implantation and expected delivery may vary with the strain.

Table 12.2  Approximate Dosing and Breeding Schedule Involved in a Two-generation Study of Effects on the Reproduction Process in Rats Age of animals (weeks) 5–9

F0/P1 Start of study Exposure for 10 weeks in diet (or other route, based on the most likely human exposure scenario)

15–19

Mating (conducted over 2 weeks or 3 estrous cycles; re-mating with a different partner is not undertaken) Gestation (approximately 3 weeks) Parturition → F1a*

21–25

Weaning (approximately 3 weeks) Growth (approximately 15 weeks)

15–19

Mating Gestation (approximately 3 weeks) Parturition → F2a* Weaning

* In certain instances, such as poor reproductive performance in controls, or in the event of treatment-related alterations in litter size, the adults may be remated to produce an F1b or F2b litter. If production of a second litter is deemed necessary in either generation the dams should be remated approximately 1–2 weeks following weaning of the last F1a or F2a litter.

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i.e. OPPTS 870.6300 (U.S. EPA, 1999) and OECD GL 426 (OECD, 2007). In this study, adverse effects of pre- and postnatal exposure on the development and function of the nervous system are characterized. Offspring are randomly selected from control and treated litters for evaluations of gross neurological and behavioral abnormalities during postnatal development and as young adults. Physical development, behavioral ontogeny, motor activity, motor and sensory function, learning and memory, and post-mortem evaluation of brain weights and neuropathology are evaluated. The history and validation of this guideline are discussed at length by Makris et al. (2009). Information on study conduct and data interpretation from a regulatory perspective can be found in OECD (2008). To ensure that children’s health risks are being adequately addressed in the risk assessment process, three areas of concern, namely, life stages evaluated, endpoints assessed and duration of exposure used in various studies, are receiving attention. Findings from studies focusing on these areas may be used to support the application of additional safety factors for specific chemicals. Developmental immunotoxicity, carcinogenesis, specialized neurotoxicity studies, endocrine disruptor studies, pharmacokinetics and direct dosing of neonates have additionally been identified as issues to be considered in the risk assessment of chemicals (Kimmel and Makris, 2001). Additional information on testing protocols, study design and interpretation of studies submitted to regulatory agencies are available in guidance documents, previous editions of this book and elsewhere (Iyer, 2001; Iyer et al., 2002; U.S. EPA, 1991, 1998a,b,c).

12.4.3  Choice of Species in Testing The laboratory species typically used to test for developmental toxicity or for reproductive effects is the rat. Some strains are considered less suitable and the rationale for the strain may vary with the compound and the effects it may cause in the species tested. The rabbit is the other species that is used as it is the one species (unlike the rat) that showed some signs for the compound thalidomide, the chemical that appeared safe in all other species tested. Additionally, because of species specificity of teratogenic agents, the exact effects noted in laboratory animals are not necessarily those observed in humans. However, all proven human teratogens have parallel but imperfect animal models. Determining which species is the most appropriate for extrapolation to humans for a given compound is difficult. Pesticides that involve food use are likely to have a higher potential exposure and are to be tested in two species as per FIFRA regulations. Among the species used for testing, the rat and mouse most successfully model the human reaction, but the rabbit is less likely than other species to give a false positive finding (U.S. EPA, 1991). The concomitant use of the rabbit with either the mouse or rat is believed to enhance the predictive potential of the individual animal model (Schardein and Keller, 1989). Accordingly, the rat and rabbit

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are commonly used. While no single species has clearly distinguished itself as being more advantageous in the detection of human teratogens over any other, it is concluded that safety decisions should be based on all reproductive and developmental toxicity data in light of the agent’s known pharmacokinetic, metabolic and toxicologic profile.

12.4.4  Choice of Testing Doses While reviewing studies submitted for regulatory purposes a number of factors are taken into consideration. A compound may be embryolethal without being teratogenic, and alternatively it can also be both embryolethal and teratogenic. The teratogenic dose range, i.e. the margin between the dose which will kill the fetus and that which does not have adverse effects on the fetus, is often very narrow. The wider the margin, the more potentially dangerous the compound is from a teratological perspective. Hence it is recommended that the highest dose should only cause slight toxic effects on the pregnant animal (e.g. decreased body weight gain) such that a majority of the pregnancies reach term (Wilson, 1979). Choice of dosing regimens is critical to determining the potential of the compound to exert adverse effects. The mid-dose must not be much lower than the high dose as such a bracketing will result in a low No Observed Adverse Effect Level (NOAEL) and not provide data on the true nature of the chemical being evaluated. While a low NOAEL may appear to be more health protective than a higher NOAEL, it is possible to miss the effects that the compound can cause at levels below the maximum tolerated dose. The choice of dose levels is critical to study design. Studies submitted with inappropriate doses are often unacceptable to regulatory agencies and the registrant has to either conduct a new study or provide justification to support the choice of doses employed.

12.4.5  Interpreting Effects Death of the conceptus may preclude expression of other major manifestations of developmental toxicity, i.e. structural abnormalities, altered growth and functional deficit. Generally the term teratogenicity is used to refer to the observation of malformations, i.e. permanent structural changes that may adversely affect survival, development, or function. Other developmental effects include variations, a term used to describe changes in fetuses that involve a divergence beyond the usual range of structural constitiution that may not adversely affect survival or health. It is sometimes difficult to distinguish between malformations and variations since the responses constitute a continuum from normal to the extremely deviant. Other terms used are anomalies, deformations or aberrations, however they are not defined any better (U.S. EPA, 1991). To further confuse this already complex issue, these other terms may get used for either of these two categories requiring a closer examination of the interpretation.

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

Evaluating variations and interpreting their incidence is illustrated in the case of supernumerary ribs (SNR). These are a common variant in some strains of mice used in standard teratology bioassays and increased incidence of SNR may be induced by a wide variety of xenobiotics and/or general maternal stress. The significance of this defect in cross-species extrapolations has been problematic. In one study in mice it has been demonstrated that SNR have a bimodal distribution composed of “rudimentary ribs” (RR) with a mode of 0.3–0.4 mm and “extra ribs” (ER) with a mode of 0.9–1.1 mm. ER and RR were foun d to be morphologically distinct; the ER were flat ended and distally joined by a cartilaginous portion, while RR were usually rounded distally and were without cartilaginous extensions. The 13th ribs were significantly longer in fetuses having SNR than in those not having SNR, whether treated or untreated. This relationship was present in all fetal ages examined and with both ER and RR groups suggesting that SNR are indicative of basic alterations in the development of the axial skeleton (Branch et al., 1996). In the case of developmental toxicity, studies are reviewed taking into account the maternal effects observed. Developmental toxic effects in the presence of severe maternal toxicity are considered less severe than those observed in the absence of maternal effects. Generally, in order to determine whether or not the conceptus is uniquely susceptible, the developmental and maternal NOAEL values are compared. The A/D ratio (Adult NOAEL:Developmental NOAEL) has been advocated previously as an index of comparative teratogenic hazard (Johnson, 1981) and has been used to characterize the developmental effects of chemicals. The strategy of carefully characterizing the observed maternal toxicity at the individual level is also employed. To determine if the malformations observed were the result of maternal toxicity, two approaches may be adopted: consideration of individual vs. group mean data and examination of data during the specific period of gestation when the developmental malformations were likely to have occurred. Furthermore, since mere correlation of maternal toxicity with fetal effects does not imply causality (Chernoff et al., 1987), maternal influence may not necessarily be the underlying mechanism of action. The severity of the effect on the fetus also needs to be considered, i.e. the effect may be severe/life threatening while the maternal effects such as slight weight loss are minor or transient. Another confounding factor is that maternal effects may be reversible whereas effects on the developing fetus may be permanent, underscoring the importance of characterization of the maternal effects. Examining the data in this manner leads to a more exacting interpretation of teratogenic potential. The use of a weight-of-evidence approach includes: Dose response. Supportive evidence in another species or related compounds. l Closer scrutiny, focusing on individual data during the discrete time period(s) in which particular fetal l l

389

malformations were most likely to have occurred, may help verify the apparent maternal toxicity at the individual level and determine the developmental toxic potential of a chemical (Iyer et al., 1999). In conducting the risk assessment the developmental toxicity study serves as a surrogate for an acute toxicity study based on the premise that the effects noted may be the result of a single exposure or a exposure over a short period. However, if the effects responsible for the NOAEL or reference dose (RFD) are known to result from multiple exposures then use of the developmental toxicity study for acute effects would be inappropriate. Similar approaches are recommended for the reproduction (two-generation) and developmental neurotoxicity studies. Effects on lactation, acceptance of offspring and sexual maturation are examined in the reproduction studies. Multi-generation studies are evaluated to determine if the effect noted is exacerbated in subsequent generations. Furthermore, spontaneous occurrence in control animals of stillborn pups and other developmental effects necessitate that evaluation of the data be subjected to rigorous statistical procedures.

12.4.6  Statistical Evaluation One of the most important aspects of developmental toxicity analysis is that the litter is to be considered the experimental unit (U.S. EPA, 1991). Since it is the maternal unit that is exposed to the compound, the effects of the test substance on each fetus in a litter are related to the status of the animal bearing that fetus. Individual differences in maternal susceptibilty can affect an entire litter, while others in the same dose group are unaffected. Hence, all fetuses in a single dose group are not equally at risk to the potential developmental effects of the test substance. Therefore the accepted practice is to consider the litter as the experimental unit for developmental toxicity studies (Collins et al., 1999; Gad and Weil, 1986; Gaylor, 1978; U.S. EPA, 1991). Evaluating overall fetal effects helps further characterize the developmental toxic effects, but the litter is the preferred unit to evaluate the effects of the compound. The use of model fitting techniques and employing the use of benchmark dose is encouraged if it is appropriate. Concurrent controls are the group of choice for comparison. Historical controls are recommended to serve as supportive evidence.

12.4.7  Exposure Assessment In evaluating the exposure for a given chemical, regulatory agencies take into consideration amount used, usage pattern (seasonality etc.) and attempt to obtain the dose that reaches either the parent’s germ cells or the developing conceptus. While developmental toxicity is usually thought to be associated with maternal/embryonic exposure, there is increasing evidence for developmental effects due to

390

male exposure (Colie, 1991; Sever, 1995). Agents associated with spontaneous abortions may also cause congenital malformations with the appropriate timing and dose, hence exposure pattern may determine the continuum of effects that might result. Additionally it is thought that the steady accumulation of pesticides in the adipose tissues during a woman’s lifetime may pose a risk, especially in the case of endocrine disrupting chemicals (Garcia-Rodriguez et al., 1996). The issues associated with the importance of timing and assessment of exposure during pregnancy have been discussed extensively (Hertz-Picciotto et al., 1996). Knowledge of the time-window of vulnerability has important implications for assessment of risks. Along with the active ingredient, organic solvents are used extensively in pesticide formulation, and hence mixers and loaders of pesticides may be exposed to higher levels of both the active ingredient as well as the solvents/inerts. Effects caused by solvents could confound the issue and may impact exposure in an adverse manner.

12.4.8  Impact of FQPA on Developmental and Reproductive Toxic Effects of Pesticides In 1996, Congress passed landmark pesticide food safety legislation supported by the administration and a broad coalition of environmental, public health, agricultural and industry groups. The bill was signed by President Clinton on August 3, 1996, and the Food Quality Protection Act of 1996 became law (P.L. 104–170, formerly known as H.R. 1627). EPA regulates pesticides under two major federal statutes. Under the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA), EPA registers pesticides for use in the United States and prescribes labeling and other regulatory requirements to prevent unreasonable adverse effects on health or the environment. Under the Federal Food, Drug, and Cosmetic Act (FFDCA), EPA establishes tolerances (maximum legally permissible levels) for pesticide residues in food. The Food Quality Protection Act (FQPA) amendments to the FFDCA direct EPA to consider a number of factors in making risk assessments as part of the tolerance setting procedure. Most of these provisions originated in recommendations from the National Academy of Sciences (NAS) 1993 report “Pesticides in the Diets of Infants and Children” and reflect concerns that children may be especially susceptible to pesticide exposure. Specifically, in setting a tolerance for pesticide residues in food, FQPA directs EPA to consider: use of an extra 10-fold safety factor to account for susceptibility of children; the special susceptibility of children, including effects of in utero exposure; cumulative effects of exposure to the pesticide and substances having a common mode of action; aggregate exposure for all consumers (i.e. other routes, such as drinking water and home and garden applications); and potential for endocrine disrupting effects. Incorporating

Hayes’ Handbook of Pesticide Toxicology

these factors into the tolerance setting process poses significant challenges to the Agency since there are many scientific uncertainties surrounding the use of these factors in risk assessment. To this end, specific areas are being brought to the Scientific Advisory Panel (SAP) for review. Current practice is generally to use a 100-fold safety factor when the toxicity data are from animal studies, and to apply extra factors of threefold to 10-fold only when specific case-by-case evidence seems to warrant it. Following extensive internal U.S. EPA discussion and deliberation, as well as the review and consideration of interim position papers (e.g. U.S. EPA, 1996b) by several FIFRA Scientific Advisory Panels, OPP released a final position paper on the application of the FQPA factor in tolerance-setting activities (U.S. EPA, 2002). The 1996 law represents a major breakthrough, amending both major pesticide laws to establish a more consistent, protective regulatory scheme. It mandates a single, health-based standard for all pesticides in all foods; provides special protections for infants and children; expedites approval of safer pesticides; creates incentives for the development and maintenance of effective crop protection tools for American farmers; and requires periodic reevaluation of pesticide registrations and tolerances to ensure that the scientific data supporting pesticide registrations will remain up to date in the future. At the current time, additional information may be accessed at the site www.epa. gov/oppfead1/fqpa/sciissue.htm. The aim of this chapter is to provide the reader with available information on the reproductive and developmental toxicity of agrochemicals used worldwide. Accordingly, a comprehensive search was conducted. This included a comprehensive review of the open literature searching databases such as Medline (PubMed) and Tox line, several secondary sources such as the Hazardous Substances Data Bank (HSDB) of the National Library of Medicine, the Integrated Risk Information System (IRIS) of the U.S. EPA, the Reprotox database, the Registry of Toxic Effects of Chemical Substances (RTECS) from the National Institute of Occupational Safety and Health, the Catalog of Teratogenic Agents (9th edition), a database compiled by Dr. T. Shepard, the Teratogen Information System (TERIS) from the University of Washington (1995), and Pesticide Information Profiles from EXTOXNET, a cooperative effort of University of California-Davis, Oregon State University, Michigan State University, Cornell University, and the University of Idaho. Animal data on the reproductive and developmental effects of agrochemicals generated for the purpose of pesticide registration have hitherto not been very accessable outside the regulatory agencies and/or the registrants producing the individual compounds. The information from these databases may be requested via the Freedom of Information Act and were searched. One such database from the California Department of Pesticide Regulation (CDPR) which may now be accessed via the internet was also searched for available online toxicological summaries

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

at http://www.cdpr.ca.gov/docs/toxsums/toxsumlist.htm. The California Department of Pesticide Regulation (CDPR) has also conducted risk assessments of compounds considered high priority based on low NOAELs for acute toxicity and possible adverse effects identified in genotoxicity studies and oncogenicity studies submitted under the Birth Defect Prevention Act (SB 950). The findings from the studies evaluating developmental toxicity and toxicity to the reproductive system for these compounds have also been reviewed and are included in this chapter. Additional reference sources include the document series such as Toxicological Profiles from the Agency for Toxic Substances and Disease Registry (ATSDR), Health Assessment Documents (HAD) and Reregistration Eligibility Documents (RED) from U.S. EPA, Monographs on the Evaluation of Carcinogenic Risk to Humans from the International Agency for Research on Cancer (IARC) and Environmental Health Criteria from the International Programme on Chemical Safety (IPCS) of the World Health Organization. Textbooks in the field were also consulted and have been listed in the references.

12.5  Toxicology studies Pesticides can be broadly classified into different categories, based on the target of use. The major classes that will be discussed include herbicides, insecticides (includes insect growth regulators), fungicides, and rodenticides. Also certain compounds that have pesticidal action (antiparasitic) are used as animal health products and will be discussed under a miscellaneous category.

12.5.1  Herbicides Included under this class of chemicals are the chlorphenoxy compounds, bipyridyls, dinitrophenols, triazines, substituted ureas, some of the carbamates, plant growth inhibitors, and amides. A number of herbicides have been studied for developmentally toxic effects and adverse effects on reproduction in animals (Table 12.3).

12.5.1.1  Chlorphenoxy Compounds Phenoxy defoliants, 2,4-D and 2,4,5-T have been used worldwide in forestry and agriculture. The phenoxy herbicide Agent Orange, composed of equal parts of 2,4-D and 2,4,5-T, received a lot of attention after its large-scale use in Vietnam by the U.S. military during the war years of 1962–1971. The phenoxy herbicide 2,4,5-T has not been manufactured since 1983. The contamination of these compounds with the dioxin TCDD (2,3,7,8-tetrachlorodibenzo-p-dioxin) in commercial preparations has further confounded the effects noted. TCDD is found to be extremely teratogenic in laboratory animals: it has a very low teratogenic minimal effective dose of 1–10 g/kg

391

in susceptible strains of mice (Pearn, 1985) and effects such as hydronephrosis and cleft palate have been noted (Couture et al., 1990). While results from Agent Orange exposure are largely inconclusive, the litigation history may be obtained in previous reports (Schuck, 1986). An increased risk of circulatory/respiratory defects, musculoskeletal/integument defects, and poly-/syn-/adactyly was seen in high-wheat verses low-wheat regions among 43,634 births, where utilized wheat acreage served as a surrogate for chlorophenoxy herbicide exposure in a record linkage study (Schreinemachers, 2003). Male offspring in high-wheat regions were almost twice as likely to have circulatory/respiratory defects compared to male offspring of low-wheat regions. And female offspring had an elevated risk of musculoskeletal/integumental defects. In the Ontario Farm Family Health Study exposure to phenoxy herbicides during the first trimester was generally not associated with increased risk of spontaneous abortion (Arbuckle et al., 1999), however, a possible role of preconception (possibly paternal) exposures to phenoxy herbicides in the risk of early spontaneous abortions is suggested. Increases in risk of early abortions (12 weeks gestation) were seen for women living on farms that used phenoxy acetic acid herbicides in the 3 months prior to conception (Arbuckle et al., 2001). In Minnesota, spring miscarriages were elevated among the spouses of male pesticide applicators who had applied one or more of the following herbicides: sulfonylurea, imidizolinone, or a mixture consisting of chlorophenoxy herbicides, sulfonylurea, and benzothiodiazole (Garry et al., 2002b).

12.5.1.2  2,4-D A recent risk assessment by Health Canada’s Pest Management Regulatory Agency concluded that the use of 2,4-D to treat lawns and turf does not entail an unacceptable risk of harm to human health or the environment (Health Canada, 2008). In considering the effects on fetal growth, no effect was seen on birth weight (Dabrowski et al., 2003), preterm birth (Savitz et al., 1997), or being small for gestational age (Savitz et al., 1997) among parents exposed to phenoxyacetic acid derivatives or fathers exposed to phenoxy herbicides. One case of cephalic malformations and severe mental retardation was noted in an infant whose parents received prolonged exposure via the dermal route from forest spraying (Casey and Colie, 1984). Though no birth defects were found, an increase in spontaneous abortions and premature births was noted in a casecontrol study examining the effects of 2,4-D (Carmelli et al., 1981). Possible adverse effects such as gestational and neonatal losses with NOAELs of 20 mg/kg/day (rats) and 500 ppm (dogs) were noted in two studies (CDPR, 2009). Markedly reduced gestational and neonatal survival was not accompanied by a commensurate degree of maternal toxicity.

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392

Table 12.3  Developmental and Reproductive Toxicity Profile of Herbicidesa Chemical

Species

Toxicity profile

Dose (mg/kg)b

Comments

References

Acrolein

Rat

Reproduction

3 (pups) 1 (parents)

Decreased body weight in pups

CDPR worksheet (1994) (Hoberman, 1991)

Alachlor (ethane sulfonate)

Rat

–

1000

Heydens et al. (1996)

Ametryn

Rat

Increased skeletal variants; Reproduction

50

Infurna et al. (1987)

Rabbit

–

60

Amitraz

Rat Mouse

Reproduction Altered estrous cycles

10.5 ppm

Pup mortality

CDPR Toxicology Summary DPN #287 (1994)

Amitrole

Rat

Thyroid effects

0.0004%

Via drinking water 3-G study

Rabbit

Abortions, reduced weight gain

4

Shalette et al. (1963) Gaines et al. (1973) CDPR worksheet 033 45711

Arsenic Arsenic acid Sodium arsenate

Teratogenic: malformations

Mouse Rat/mouse

Asulam

Rat

Reproduction decreased 1000 ppm number of live births

Atrazine

Rat

Fetal toxicity

7.5

70

Parenterally (i.p)

CDPR: DPN # 180 Toxicology Summary (1992) Beaudoin (1974) Hood and Bishop (1972) CDPR worksheet (360) 010 25257; 19

2-G study

Disruption of ovarian cycle, induced pseudopregnancy

Infurna et al. (1986) Cooper et al. (1996)

Reproduction

Balagrin

Clopyralid

Rabbit

Fetal toxicity

75

Infurna et al. (1986)

Mouse

Teratogenic

22

Ivanova-Chemishanksa et al. (1979) [2]

Rat

Teratogenic

1/50 LD50

Mirkova (1980)

Rat

Delayed ossification

75

Reproduction

Selectively toxic

Hayes et al. (1984)

2-G study

Dietz et al. (1986) in CDPR 2009 database

Rabbit

–

250

Cyanazine

Rat

Developmentally toxic

1

2,4-D

Mouse

Developmentally toxic, teratogenic

221

Courtney et al. (1977)

Rat

Fetal death Teratogenic Reproduction

50

Khera and McKinley (1972)

Hamster

Teratogenic

20

Mouse

Developmentally toxic Teratogenic

0.2%

2,4-D  picloram

Hayes et al. (1984) Anophthalmia/ microphthalmia

Lu et al. (1982)

2-G study Collins and Williams (1971) Drinking water route

Blakely et al. (1989)

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

393

Table 12.3  (Continued) Chemical

Species

Toxicity profile

Dose (mg/kg)b

2,4-D  2,4,5-T

Rat

Behavioral effects

50/125

Mohammad and St Omer (1988)

2,4-D butoxyethanol Rat ester Rabbit

Teratogenic

150

Khera and McKinley (1972)

–

75

Liberacki et al. (1994)

2,4-D butylester

Rat

Teratogenic

150

Khera and McKinley (1972)

Sheep

Teratogenic Sterility, fetal death

3 kg 3%/10-ha pasture

Sadykov et al. (1972) [1]

2,4-D diethylamine

Rat

Teratogenic

0.5 LD50

Aleksashina et al. (1973)

Bifenox

Mouse

–

100

Francis (1986)

Bromacil

Rat

–

250 ppm

Diet

Rabbit

–

250 ppm

Diet

Mouse

Developmentally toxic

96.4

Rat

Skeletal variation

15

Butiphos

Rat

Developmentally toxic

12.5

Kasymova (1975) [1]

Buturon

Mouse

Increased mortality Teratogenic

100

Matthiaschk and Roll (1977) [1]

Chloramben

Rabbit

Intrauterine growth retardation

500

CDPR worksheet (266) 011 36993; 1987

Chloridazon

Rat

Resorption

1/50 LD50

Dinerman et al. (1970) [1]

4-Chloro-2-methyl phenoxyacetic acid ethyl ester

Rat

Teratogenic

100

Yasuda and Maeda (1972) [1]

2-Chlorophenyl-4nitrophenyl ether

Mouse

–

1000

Francis (1990)

Chloroprophan

Mouse

Developmentally toxic Teratogenic

3000 750

Tanaka et al. (1997)

2,4-D isooctyl ester Rat

Teratogenic

150

Khera and McKinley (1972)

2,4-D isopropylamine

Rabbit

–

75

Liberacki et al. (1994)

2,4-D propylene glycol butyl ether ester

Rat

Developmentally toxic

87.5

Schwetz et al. (1971) [1]

2,4-D Rabbit triisopropanolamine

–

75

Liberacki et al. (1994) [2]

Dalapon

Rat

Skeletal effects

500

CDPR worksheet 006 036526; 1986

Daminozide

Rat

–

1000

Khera et al. (1979b)

Diallate

Rabbit

–

10

Johaunsen et al. (1977) [1]

Bromoxynil

Comments

References

Sherman (1968) [1]

Rogers et al. (1991)

(Continued )

Hayes’ Handbook of Pesticide Toxicology

394

Table 12.3  (Continued) Chemical

Species

Toxicity profile

Dose (mg/kg)b

2,4-DM

Rat

Developmentally toxic

3.4

Sokolova (1976) [1]

2,5-Dichlorophenyl Mouse 4-nitrophenyl ether

–

1000

Francis (1990)

3,4-Dichlorophenyl Mouse 4-nitrophenyl ether

–

1000

Francis (1990)

2,6-Dichlorophenyl Mouse 4-nitrophenyl ether

–

500

Francis (1990)

2,5-Dichlorophenyl Mouse 4-nitrophenyl ether

–

1000

Francis (1990)

2,3-Dichlorophenyl Mouse 4-nitrophenyl ether

–

400

Francis (1990)

Dichloroprop

Mouse

Teratogenic

400

Roll and Matthiaschk (1983)

Rat

Postnatal behavioral effects

5

Buschmann et al. (1986)

Dicotex

Rat

Teratogenic

20

Gzhegotskii and Shtabskii (1969) [1]

Dicuran

Rat

–

5000

[1]

Dinoseb

Rat

Teratogenic Reproductive system

200 ppm

Diuron

Rat

–

500

Khera et al. (1979a,c)

Endothall

Rat

–

25

Trutter et al. (1995)

EPTC

Rat

–

1/20 ��� LD50

Ethalfluralin

Rat

–

1000

Rabbit

–

300

Hamster

Teratogenic

20

Rat

–

400

Rabbit

–

40

Ethofumesate

Rabbit

Increased resorptions, delayed ossification

30

CDPR Toxicology Summary (1993)

Fluoxypyrmethylheptyl ester

Rat

Skeletal variations

600

Carney et al. (1995)

Hexachlorobutadiene

Rat

Reduced fetal weights Reproductive system

15 ppm

Inhalation route (fertility study)

Saillenfait et al. (1989) Schwetz et al. (1977) [1]

Hexazinone

Rat

Reproductive system

5000 ppm

Diet; 3-G study

Kennedy and Kaplan (1984)

Rabbit

–

125 ppm

Mouse

–

100 ppm

Ethephon

Ioxynil octanoate

Comments

References

Diet, only Giavini et al. (1986) developmentally Hall et al. (1978) [1] toxic by p.o. route

1/20 LD50

Medved et al. (1970) [1] Byrd et al. (1990)

Minta and Biernacki (1981)

Kobayaski et al. (1976) [1]

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

395

Chemical

Species

Toxicity profile

Lenacil

Dog

–

Rat

Reproductive system

Rat

Teratogenic Reproductive system Reproductive system Male reproductive system development

200 100 ppm

3-G study

12.5

2-G study

Rabbit

–

125 ppm

Diet

Maleic hydrazide

Rat

–

1600 ppm

MCPA

Rat

Embryotoxic Teratogenic

1/2 LD50 1/2 LD50

Mouse

Teratogenic

200

Mouse

Teratogenic

400

Rat

Postnatal toxicity

13

Meturin

Rat

–

{1/ 10} LD50

Sadovskii et al. (1976) [1]

Molinate

Rat

Teratogenic, Increased resorptions, Intrauterine growth retardation

35

CDPR Toxicology Summary, 1998, DPN # 228

Rat

Reproduction: sperm 5 ppm males abnormalities, detached 20 ppm females heads, ovarian interstitial tissue vacuolation

Linuron

Mecoprop

500 ppm

–

1000

Monolinuran

Mortality Teratogenic

25 25

Naphoxyacetic acid Rat

–

250

Nitrofen

Mouse

Growth retardation Teratogenic

250

Rat

Teratogenic Reproductive system

121

Hamster

Teratogenic

400

Hamster

–

2000

Rat

Reproductive system

Rat

Depressed fetal weight

Rabbit

–

125

Rat

Teratogenic, Postimplation loss, incomplete ossification

12

Rabbit

Teratogenic, Resorptions, 60 constraint-related arthrogryposis

Norea

Oryzalin

Oxadiazon

References

Diet

Worden et al. (1974) [1]

3-G study

3-Monochlorophenyl Mouse 4-nitrophenyl ether Mouse

Comments

Hodge et al. (1968) [1]

Matthiaschk and Roll (1977) [1]

Dermal/oral 3-G and fertility Studies

Schardein (1993) 3-G study

Kennedy et al. (1973) [1] Dickson (1979)

CDPR Toxicology Summary (1995), DPN # 346

(Continued )

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396

Chemical

Species

Toxicity profile

Comments

Oxyfluorfen

Rat

Teratogenic, early resorptions, decreased fetal weight, skeltal malformations and variations

15

Paraquat

Mouse

–

100

References CDPR Toxicology Summary (1993), DPN # 381

Reproductive system

2-G study

1,2,3,7,8-PentaMouse bromodibenzofuran

Teratogenic

4000

1,3,4,7,8-PentaMouse bromodibenzofuran

Teratogenic

2400

Phosphinothricin

Rat

Fetotoxicity

250

Rabbit

Fetotoxicity

20

Rat

–

1000

Rabbit

–

400

Rabbit

–

500

Pichloram Rabbit triisopropanolamine

Abortion

1000

Pometryn

Rat

Teratogenic

25

Gzhegotskii et al. (1970) [1]

Propachlor

Rat

Equivocally teratogenic

1/5 LD50

Mirkova (1975) [1]

Propanil

Rat

Reproductive system

Propazine

Rat

Decreased fetal weight

Siduron

Rat

Reproductive system

Pichloram

Pichloram ethylhexyl ester

3-G study 1/5 LD50

Dinerman et al. (1970) [2] 3-G study 3

Sherman (1971) [1]

Simazine

Rat

Teratogenic

0.2 mg/m

SLA 3992

Rat

Teratogenic

20

Rabbit

Teratogenic

20

Mouse

Teratogenic

15

Rat

Teratogenic Reproductive system

50

Rabbit

–

40

Hamster

Teratogenic

20

Sheep

–

113

Primate

Growth retardation Abortion

40

Rat

Teratogenic

50

Sokolik (1973) [1]

Mouse

Teratogenic

74

Newbert and Dillman (1972) [1]

Mouse

–

9

Newbert and Dillman (1972) [1]

2,4,5-T

2,4,5-T butyl ester

2,4,5-T phenol

Inhalation route

Gaines et al. (1975) [1] 3-G study Thompson et al. (1971) [1]

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

Chemical

Species

Toxicity profile

2,4,5-T propylene glycol butyl ether ester

Sheep

–

100

Tebuthiuton

Rat

–

1800 ppm

Triallate

Rabbit

–

10

Trichloroacetic acid Rat

Developmentally toxic Teratogenic

330 330

Trichloropyr butoxyethyl ester

Rat

Developmentally toxic

300

Trichloropyr triethylamine

Rat

Developmentally toxic

300

Tridiphane

Mouse

Resorption Teratogenic

250

Rat

Skeletal variants Reproductive system

100

Mouse

Skeletal variation

1

Rat

Depressed fetal weight

1000

Rabbit

Developmentally toxic

500

–

1000

Trifluralin

Triisopropanolamine Rat

Of other phenoxy herbicides that have been studied, 4-chloro-2 methylphenoxy acetic acid ethyl ester caused (31% incidence) cleft palate, and anomalies of the heart and kidney in rats (Schardein, 1993). 4-chloro-o-toloxy acetic acid also was found to be teratogenic in mice and rats at high oral doses (Roll and Matthiaschk, 1983; Schardein, 1993).

12.5.1.3  Amitrole This nonselective postemergence herbicide is also an antithyroid agent. While structural malformations were not noted when tested in animal studies, fetal thyroid lesions were observed in rats exposed to amitrole via drinking water (Schardein, 1993). Given the role of the fetal thyroid in neurological development, this pesticide may affect

397

Comments

References

Diet

Todd et al. (1974) [1] Johannsen et al. (1977) [1]

Hanley et al. (1996) [1]

(2-G and repro study)

cognitive abilities of individuals subsequent to in utero or perinatal exposure.

12.5.1.4  Bromoxynil The developmental toxicity of the wide-spectrum herbicide bromoxynil (bromoxynil phenol; 3,5-dibromo-4-hydroxyphenyl cyanide) and its octanoate ester (2,6-dibromo-4cyanophenyl octanoate) were evaluated in Sprague–Dawley rats and Swiss–Webster mice. Highest doses of both compounds increased the incidence of supernumerary ribs (SNR) in fetuses of treated rats, but did not induce other anomalies (Rogers et al., 1991). In the teratogenicity study submitted to CDPR, rats were dermally exposed to Buctril (containing 33.8% bromoxynil octanoate) diluted with water. Based on a dose-dependent increase in the incidence

398

of extrathoracic ribs in fetuses, at the 15 mg/kg/day level and above, the developmental NOAEL was determined to be 10 mg/kg/day. Recent data suggest that bromoxynil as well as other chemicals such as retinoic acid and valproic acid may affect developmental processes involved in the patterning of the axial skeleton (Kawanishi et al., 2003).

12.5.1.5  Dinoseb Dinoseb (2-sec-butyl-4,6-dinitrophenol) has been shown to produce substantial spermatotoxicity after 1 to 5 doses in short duration tests (Linder et al., 1992). In mice at 17.7 mg/kg/day, subcutaneous or intraperitoneal administration of dinoseb during organogenesis resulted in skeletal defects, cleft palate, hydrocephalus, and adrenal agenesis. Maternal toxicity was however noted at doses between 17.7 and 20 mg/kg/day (Gibson, 1973). Also dinoseb has been reported to produce a high incidence of dilated renal pelvis in the term rat fetus (McCormack et al., 1980) as well as supernumerary ribs (SNR) in mice (Kavlock et al., 1985). Teratogenic effects such as increased incidence of microphthalmia were also reported in the rat fed dinoseb in the diet (Giavini et al., 1986). Eye defects and neural malformations were noted in the rabbit leading to its banning by the U.S. EPA in 1986. In a recent review, the developmental toxicity of dinoseb was noted to be remarkably different between animal species and varied with the route of exposure (Matsumoto et al., 2008).

12.5.1.6  Bipyridyl Compounds   (Paraquat, Diquat) The herbicide paraquat has resulted in at least eight fetal deaths when taken during pregnancy as a result of maternal poisoning (Talbot et al., 1988). However no adverse effects were reported in animal developmental and reproductive toxicity studies submitted to CDPR. On the other hand, for the herbicide diquat, adverse systemic effects were noted in the rat in parents and offspring (cataracts and eye pathologies in both sexes of F0 and F1 at  240 ppm; an increase of hypertrophy and hyperplasia of collecting duct epithelium and tubular dilatation in the renal papilla in both sexes of F1 at 240 ppm; F1 and F2 pups showed hydronephrosis at  240 ppm). In data submitted to CDPR for the rat, the developmental NOAEL  12 mg/kg/day with intrauterine growth retardation as measured by decreased weight and delayed skeletal ossification, and hemorrhagic kidneys as the main effects observed (CDPR, 2009). Mice appear to be more sensitive to diquat than rats. The NOAEL in the mouse for both maternal toxicity (clinical signs, death) and developmental toxicity (skeletal anomalies, exencephaly, and umbilical hernia) was 1.0 mg/kg/day. However the rabbit appeared to be the most sensitive laboratory animal to diquat in developmental toxicity studies. The NOAEL for maternal toxicity (histopathological changes in the liver,

Hayes’ Handbook of Pesticide Toxicology

intestine, and vasulature; mortality) was 3.0 mg/kg/day but the developmental NOAEL was below 1.0 mg/kg/day. Delayed ossification of the ventral tubercle of the cervical vertebrae was noted in all treatment groups compared to the controls. The incidence of fetal malformations was significantly greater in the low dose (1.0 mg/kg/day) and the high dose (10 mg/kg/day) compared to the controls; while the mid-dose lacked statistical significance it appeared biologically significant (more than a twofold increase over controls) and hence supportive of a treatment-related effect. Paraquat on the other hand has been linked with causation of Parkinson-disease-like symptoms.

12.5.1.7  Ethyl Dipropyl Thiocarbamate (EPTC) The thiocarbamate class of pesticides has been shown to cause a wide range of effects. EPTC was determined to cause adverse developmental effects in the rat with a NOAEL  30 mg/kg/ day based on increased resorptions at levels below the maternal NOAEL (100 mg/kg) (CDPR, 2009).

12.5.1.8  Ethofumesate In a developmental toxicity study in rabbits ethofumesate was determined to have adverse effects with a developmental NOAEL of 30 mg/kg/day based on increased resorptions and delayed ossification noted at higher doses (300 and 3000 mg/kg/day) tested. Maternal toxic effects such as abortions and death were noted at the high dose of 3000 mg/kg/day resulting in a maternal NOAEL of 300 mg/ kg/day (CDPR, 2009).

12.5.1.9  Molinate Results from several studies on the herbicide molinate have consistently demonstrated that exposure of male laboratory animals to the compound via the oral/inhalation route causes a decrease in fertility, abnormal sperm morphology, decreased epididymal sperm number, and/or testicular degeneration. Unexposed females mated to exposed males (rabbits/mice/rats) had significant (p  0.05) preimplantation loss, possibly a result of the inability of the sperm to fertilize ova. Female rats and mice exposed to molinate in the diet also exhibited significantly (p  0.05) reduced litter sizes, along with histopathological abnormalities in the ovaries such as vacuolation and hypertrophy of the thecal/ interstitial cells (CDPR, 2009). Recent data do suggest that humans are probably less sensitive and less likely than rats to experience the reproductive toxicity of molinate largely due to unequal rates of metabolism of molinate to molinate sulfoxide (Jewell and Miller, 1998). However, the relative degree of risk cannot be quantified at this time. In the rat, molinate demonstrated adverse effects such as increased resorptions and intrauterine growth prior to the onset of maternal toxicity with a developmental NOAEL of 35 mg/kg/day.

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

12.5.1.10  Nitrofen This pre- or postemergence herbicide induced a high incidence of diaphragmatic hernia and harderian gland alterations in mice fetuses subsequent to maternal oral exposure (Gray et al., 1983; Nakao et al., 1981). In rat studies hydronephrosis and respiratory problems were noted (Costlow and Manson, 1980) whereas eye abnormalities were noted following percutaneous exposure to the dams (Francis and Metcalf, 1982). Exposure during only two gestational days altered the development of the para- and mesonephric ducts resulting in renal malformations in females and agenesis of the vas, epididymis, and seminal vesicles in males (Gray et al., 1985). The teratogenic activity of nitrofen has been attributed to alterations in maternal and fetal thyroid hormone status (Manson et al., 1984).

12.5.1.11  Triazines This class of compounds is heavily used throughout the world. They include herbicides such as atrazine, cyanazine, propazine and simazine, and the insecticide cyromazine. Eye defects such as anophthalmia, cryptophthalmia, microphthalmia, and cyclopia have been noted in animal studies for some of these compounds (CDPR, 2009). In a reproduction study in rats exposed to cyanazine, the significant toxicological finding was decreased pup viability: F1a pups at 250 ppm on day 21 and F2a pups at 150 and 250 ppm on day 4 (NOAEL  75 ppm). A possible adverse effect was indicated (pup NOAEL  adult NOAEL). In the rat teratogenicity study, the developmental toxicity NOAEL was 5 mg/kg/day (increased number of fetuses and pups with microphthalmia or anophthalmia at 25 and 75 mg, decreased litter size and weight at 75 mg, increased total litter resorptions at 25 and 75 mg/kg, and decreased live litter size and survival to day 21 of lactation at 75 mg). Since developmental toxicity was seen at levels of cyanazine causing only slight maternal toxicity, the effects were considered adverse (CDPR, 2009; Iyer et al., 1999).

12.5.1.12  Urea Herbicides Several urea herbicides induce genetic abnormalities in standard tests for genotoxic potential. They are generally the phenylureas and the effects of some of these compounds are detailed below. Diuron, a widely used substituted urea herbicide, induced wavy ribs at doses of 250 and 500 mg/kg/day (mid and high dose) in rats. Ossification of the calvarium was delayed in fetuses of dams that received 125 mg/kg with the study yielding no NOAEL (Khera et al., 1979c). However diuron did not produce any adverse effects for either reproduction or teratogenicity in studies submitted for registration (CDPR, 2009). Reproductive abnormalities, particularly those affecting sperm morphology and function, were noted in rats

399

exposed to isoproturon (Behera and Bhunya, 1990). Maturational malformation of sperm and retarded spermatogenesis were also observed (Sarkar et al., 1997). Linuron, an antiandrogenic pesticide, has been shown to induce a level of external effects consistent with its low affinity for the androgen receptor (AR) resulting in reduced anogenital distance, retained nipples, and a low incidence of hypospadias as well as malformed epididymides and testis atrophy (Gray et al., 1999b). Additionally, linuron may produce Leydig cell tumors via an antiandrogenic mechanism where sustained hypersecretion of luteinizing hormone (LH) appears to be responsible for the development of Leydig cell hyperplasia and adenomas (Cook et al., 1993). Linuron may display several mechanisms of endocrine toxicity, one of which involves AR binding (Gray et al., 1999b). Linuron produced malformations in the rat at 100 mg/kg/day, but did not demonstrate teratogenic potential in the rabbit. In a two-generation reproduction study in rats an effect on overall development (statistically significant decrease in pup weights, litter size, and pup viability: day 0–4) was detected yielding a pup NOAEL of 100 ppm in diet equivalent to 8.3 mg/kg/day (U.S. EPA, 1995). Monolinuron, a related compound, has been shown to cause cleft palate in the mouse (Schardein, 1993).

12.5.2  Insecticides Included in this class of chemicals are the organophosphates, organochlorines, chlorinated cyclodienes, and carbamate esters. Many insecticides have demonstrated developmentally toxic effects and adverse effects on reproduction in animals (Table 12.4). Recently, reproductive toxicity of combined effects have been examined (Gomes et al., 2008). Birth outcome was studied in pre-partum litters of mice exposed to oral doses of organophosphorous pesticides at low and high concentrations before mating. Exposed and unexposed pregnant dams were delivered by caesarean section 1 day before partum, the fetuses were collected, counted and weighed, and the numbers of resorptions were recorded. Live litter sizes were nonsignificantly higher in all the exposed groups compared with the control group. The numbers of resorptions were significantly higher in all the exposed groups than in the comparison groups. The incidence of intrauterine growth retardation was significantly higher in all the exposed groups than in the comparison groups. The incidences of congenital malformations were significantly higher in the exposed groups than in one or more of the comparison groups for the defects of the ears, eyes, jaws, brain, and tongue in all the exposed groups. Low-set microtia, cataract or open eyelids, microcephaly or anencephaly, maxillary or mandibular hypoplasia, and protruding tongue were observed in all groups, but the numbers were significantly higher in the exposed groups compared with one or more of the comparison groups. Curled or missing tail and intra-auricular

Hayes’ Handbook of Pesticide Toxicology

400

Table 12.4  Developmental and Reproductive Toxicity Profile of Insecticidesa Chemical

Species

Toxicity profile

Dose (mg/kg)b

Aldicarb

Rat

Acutely toxic, hence not considered teratogenic

0.04

Rabbit Aldrin

Comments

References Risher et al. (1987)

0.1

Mouse

Teratogenic Reproduction

Rat

Reproduction

Hamster

Embryotoxic Teratogenic

Dog

Reproduction

Sheep

Teratogenic

1

Younger (1965) [1]

Rat

Equivocally teratogenic

10

Kimbrough and Gaines (1968) [1]

Bendiocarb

Rat

Reproduction

2

1,3-Bis (carbamoylthio)-2-N,Ndimethylamino propane

Mouse

–

100

Rat

–

100

Hamster

Equivocally teratogenic

100

Bromophos

Mouse

–

183

Nehez et al. (1986)

Carbaryl

Mouse

–

150 p.o. or 5660 ppm (diet)

Murray et al. (1979a) [1] DeNorsica and Lodge (1973) [1]

Apholate

Rat

Reproduction

25

(6-G study)

Ottolenghi et al. (1974) [1] Deichmann and Keplinger (1966)

3-G study

Hodge et al. (1967) [1]

50 50

Ottolenghi et al. (1974) [1] Deichmann et al. (1971)

DPN # 50094

CDPR Toxicology Summary (1997) Mizutani et al. (1971c) [1]

2000

(2-G study)

500

3-G study

Weil et al. (1972) Collins et al. (1971)

Carbofuran

Dog

Resorption, teratogenic

6.25

Smalley et al. (1969)

Pig

–

30

Smalley et al. (1969)

Hamster

Fetal mortality

125

Robens (1969) [1]

Rabbit

Teratogenic

150

Murray et al. (1979a) [1]

Guinea pig

Teratogenic

300

Robens (1969) [1]

Sheep

Teratogenic

250 ppm

Primate

Abortion

2

Dougherty (1971)

Cow

–

5.5

Macklin and Ribelin (1971) [1]

Rat

Male reproductive system effects

0.2 0.4 3

Diet

Adults In utero, lactation

Weil et al. (1972)

Pant et al. (1995, 1997) McCarthy et al. (1971) [1]

Dog

–

50 ppm

McCarthy et al. (1971) [1]

Rabbit

–

0.5

Schardein (1993)

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

401

Table 12.4  (Continued) Chemical

Species

Toxicity profile

Dose (mg/kg)b

Chlordane

Rat

Reproduction

8000

Chlordecone

Mouse

Reproductive failure

40 ppm

Rat

Fetotoxic Reproduction

10

Chlordimeform

Rat

Postnatal behavioral deficit

100 g

Chlorfenvinphos and carbaryl

Rat

Ossification disorders

1/20 LD50

Chlormequat chloride

Rat

–

1000 ppm

Hamster

Teratogenic

100

Juszkiexicz et al. (1970) [1]

Mouse

Fetotoxic

1

Deacon et al. (1980)

Rat

Reproduction

15

Ciafos

Rat

–

10

Coumaphos

Cow

–

28 g /45 kg BW

Crotoxyphos

Cow

–

3.1

Macklin and Ribelin (1971) [1]

Cyfluthrin

Rat

Reproduction; reduced pup viability; fetal malformations Postimplantation loss

50 ppm 0.46 mg/m3 20 20

CDPR Toxicology Profile DPN # 50317

Cypermethrin

Rat

Developmentally toxic Teratogenic

1/40 LD50

Shawky et al. (1984)

DDT

Mouse

Reproduction

7 ppm

Diet 6-G study

Ware and Good (1967) [1] Deichmann and Keplinger (1966)

Rat

Reproduction

1/50 LD50

3-G study

Dinerman et al. (1970) [1] Ottoboni (1969) [1]

Rabbit

Developmentally toxic

10

Hart et al. (1972) [1]

DEET

Rabbit

Incomplete ossification, other skeletal effects

30

CDPR DPN # 50191 Toxicology Summary (1999)

Deltamethrin

Rat

Teratogenic

Demethyl-bromophos sodium

Mouse

–

86

Nehez et al. (1986)

Demethyl-bromophos tetramethyl-ammonium

Mouse

–

85

Nehez et al. (1986)

Dialifor

Hamster

Teratogenic

100

Robens (1970a) [1]

Chlorpyrifos

Rabbit

Comments

References Usami et al. (1986) Ambrose et al. (1953)

Fertility study

Huber (1965) Good et al. (1965) [1] Chernoff and Rogers (1976) Cannon and Kimbrough (1979)

Diet

Olson et al. (1978) [1] Tos-Luty et al. (1974) [1]

Diet

(2-G study)

Schardein (2000)

Breslin et al. (1996b) Yamamoto et al. (1972a) [1]

Topical route

Bellows et al. (1975)

Kavlock et al. (1979)

(Continued )

Hayes’ Handbook of Pesticide Toxicology

402

Table 12.4  (Continued) Chemical

Species

Toxicity profile

Dose (mg/kg)b

Diazinon

Rat

Equivocally teratogenic Reproduction; decreased pup survival, decreased ovary weights

95–100 10 ppm

Dobbins (1967) CDPR Toxicology Summary (1999)

Rabbit

–

30

Robens (1969) [1]

Hamster

–

0.25

Mouse

Reduced postnatal growth Behavioral deficits

0.18

Spyker and Avery (1977)

Cow

–

6.6

Macklin and Ribelin (1971) [1]

m-Dichlobenzene

Rat

–

200

Ruddick et al. (1983)

p-Dichlobenzene

Rat

Reproduction

Rabbit

–

800

Dicresyl

Rat

–

4000 ppm

Diet

Yasuda (1972) [1]

Dieldrin

Mouse

Teratogenic Reproduction

15

6-G study

Ottolenghi et al. (1974) [1] Deichmann and Keplinger (1966)

Rat

– Reproduction

6

Hamster

Embryotoxic, teratogenic

30

Ottolenghi et al. (1974) [1]

Rabbit

–

6

Dix et al. (1977) [1]

Sheep

–

25 ppm

Hodge et al. (1967) [1]

Dog

–

0.2

Kitselman (1953) [1]

N,N-Diethyl benzene sulfonamide

Rat

Teratogenic

300

Leland et al. (1972)

Rabbit

Resorption

25

N,N-Diethyl-ntoluamide

Rat

Reduced fetal weight Reproductively neurotoxic

750

Schoenig et al. (1994) Wright et al. (1992)

Rabbit

–

325

Schoenig et al. (1994)

Pig

–

100 ppm

Fertility study

Escobar et al. (1980a) [1]

Sheep

–

100 ppm

Fertility study

Escobar et al. (1980a) [1]

Rat

Teratogenic Reproduction

3 3-G study

Khera et al. (1979c) Levinskas et al. (1966a) [1]

Cat

Teratogenic

12

Mouse

Reproduction

Mouse

Developmentally toxic

Diflubenzuron

Dimethoate

O,O-Dimethyl-S-(2acetylaminoethyl) dithiophosphate

Comments

2-G study

Neeper-Bradley et al. (1989) [1] Hayes et al. (1985)

3-G study

Chernoff et al. (1975) Hodge et al. (1967) [1]

Khera et al. (1979a) 5-G study

8

References

Budreau and Singh (1973b) Hashimoto et al. (1972) [1]

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

403

Chemical

Species

Toxicity profile

Comments

References

Empenthrin

Rat

–

Endosulfan

Rat

Delayed ossification Reproduction

500

Fertility study

Kaneko et al. (1992)

5

3-G study

Endrin

Mouse

Teratogenic Reproduction

2.5

Hamster

Embryotoxic Behavior deficits Teratogenic

5 5 0.75

Ethohexadiol

Rat

Developmental study

2 ml/kg

Fenamiphos

Rat

–

3

Rabbit

–

2.5

Fenbutatin oxide

Rat

Reproduction; decreased pup weight gain in F1 and F2

250 ppm

Fenitrothion

Mouse

–

80 ppm

Rat

Postnatal behavioral deficits

10

Fensulfothion

Rabbit

Teratogenicity, malformations, incomplete ossification

0.1

Fenthion

Rat

–

18

Rat

Epididymal cytoplasmic vacuolation

2 ppm

Fluvalinate

Rabbit

Teratogenicity, malformations

25

CDPR Toxicology Summary (1988) DPN # 50241

Formothion

Rabbit

–

30

Klotsche (1970) [1]

Heptachlor

Rat

Cataracts in both generations

10 ppm

Heptachlor and heptachlor epoxide

Rat

–

7 ppm

Imidazolidinone

Rat

–

240

Isobenzan

Mouse

–

1 ppm

Isofenphos

Rat

–

10

Leptophos

Rat

Developmentally toxic

12.5 ppm

Lindane

Mouse

Reduced fetal growth

Rat

Decreased fertility Mortality and developmental delay –

0.5 0.5 100 ppm

Rabbit

Inhibited development

40

Dzeirzawski (1977) [2]

Hamster

Inhibited development

20

Dzeirzawski (1977) [2]

Ottolenghi et al. (1974) [1] Good and Ware (1969) [1]

Occlusive cutaneous

CDPR Toxicology Summary, (1994) DPN # 214 Diet

CDPR worksheet 234-084 054352 (1987)

2-G study

CDPR worksheet

2-G study Eisler (1970) [1]

Diet

Ware and Good (1967) [1] Mast et al. (1985) [1]

Diet Yamagishi et al. (1972) [1] 3-G study

(Continued )

Hayes’ Handbook of Pesticide Toxicology

404

Chemical

Species

Toxicity profile

Malathion

Rat

– –

300 240

Rabbit

–

100

Rat

Teratogenicity, decreased fetal weights, severe malformations

5

Rabbit

Postimplantation loss, severe defects (cleft palate, meningiocele)

5

Rabbit

–

2.5

Rabbit

–

100 ppm

Rat

Reproduction

Methyl demeton

Rat

Reproductive toxicity

Methyl ISP

Mouse

–

2.5 mg/g BW

Rat

–

2.5 mg/g BW

Rat

Developmentally toxic

5

Teratogenic

5

Metam-sodium

Methamidaphos Methomyl

Methyl parathion

Comments 2-G study

CDPR Toxicology Summary (1994) DPN # 50150

Diet

Rat

–

100

Nicotine

Mouse

Resorption Teratogenic

0.008 g/g BW

Rat

Developmentally toxic Postnatal behavioral deficits

2

s.c. route

2

i.v. route

Rabbit

–

5

s.c. route

Cow

–

Sheep

–

Rat

– Reproduction

100

Rabbit

Reproduction

4

Mouse

–

75

Hamster

–

1.25

Rat

Embryotoxic Developmentally toxic Reproduction

4

Pentachlorophenol

Kaplan and Sherman (1977) [1]

Gofmekler and Khuriev (1971) [1]

Naled

Oxamyl

References

Hudson and Timiras (1972) [1] Navarro et al. (1989) [1] Vara and Kinnunen (1951) [1] Keeler (1980a) [1]

1 and 3 G studies

Hinkle (1973) [1] Less toxic by p.o. route

Welch et al. (1985) [1] Schwetz and Gehring (1973) [1] Schwetz et al. (1978b) [1]

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

405

Chemical

Species

Toxicity profile

Phorate

Rat

Reproduction

Mouse

Reproduction

Phosalone

Rat

–

Phosfolan

Rat

Teratogenic

Phosmet

Rat

Embryotoxic Teratogenic

0.3 0.3

Rabbit

–

35

Mouse

–

0.6

Rat

–

0.6

Rabbit

–

10

Rat

Cataracts and reduced survival in generational study

20 ppm

Mouse

Teratogenic

660

Ogata et al. (1981) [1]

Rat

–

3000

Kennedy et al. (1977) [1]

Rabbit

Equivocally teratogenic

100

Schwetz et al. (1976) [1]

Potassium arsenate

Sheep

–

0.75

James et al. (1966) [1]

Propoxur

Rat

Neonatal CNS impairment

1000 ppm

Ronnel

Rat

Developmentally toxic

2.5

Rabbit

Teratogenic

Fox

Teratogenic

100

Rotenone

Rat

Embryotoxic

2.5

Sarin

Rat

–

380 g/kg

Rabbit

–

15 g/kg

Sodium arsenite

Hamster

–

25

Sodium selenite

Hamster

Teratogenic

90

Soman

Rat

–

165 g/kg

Rabbit

–

15 g/kg

Rabbit

Neonatal mortality

0.3

Photodieldrin

Photomirex

Piperonyl butoxide

Sulfluramid

1.94 mg/m3

Comments

References

Inhalation route 3-G study

Dilley et al. (1977) [2] Levinskas et al. (1966a) [1]

3-G study

Levinskas et al. (1966a) [1]

1/10 LD50

Michailova and VachkovaPetrova (1976) [1]

Selectively toxic

Kagen et al. (1978) [1] Fabro et al. (1966) [1]

Villeneuve et al. (1978) [1] Reproduction study

Diet

Rosenstein and Chernoff (1976) [1] Khera et al. (1982)

Teratogenic by i.v. route

(Continued )

Hayes’ Handbook of Pesticide Toxicology

406

Chemical

Species

Toxicity profile

Comments

2,3,5,6-Tetrachloropyridine

Rat

–

150

Thiometon

Rabbit

–

5

Toxaphene

Mouse

Teratogenic

15

Rat

Decreased skeletal ossification

35

References

Klotzsche (1970) [1]

Reproduction

3-G study

Kennedy et al. (1973) [1]

Tribufos

Rat

–

28

Trichlorfan

Mouse

–

300

Martson (1979) [1]

Rat

Teratogenic

400

Hamster

Teratogenic

400

Staples and Goulding (1977) [1]

Pig

Teratogenic

60

Knox et al. (1978) [1]

Trichloro acetonitrile

Rat

Teratogenic

7.5

Smith et al. (1986a) [1]

1,2,3-Trichlorobenzene

Rat

–

600

1,2,4-Trichlorobenzene

Rat

–

300

1,3,5-Trichlorobenzene

Rat

–

600

Triphenyltin hydroxide

Rat

–

20

Equivocally reduced fertility

100 ppm

Teratogenic

90

Testicular toxin

0.7

Valexon

Rat

septal or intraventricular septal defects were observed in higher numbers in the groups in which both the males and the females were exposed than in the comparison groups. Male:female sex ratios were significantly higher in the groups in which males only and females only were exposed (Gomes et al., 2008).

12.5.2.1  Aldrin Prenatal aldrin exposure induced developmental changes (a decrease in the median effective time for incisor teeth

Wine et al. (1978) [1] Reproduction Study

Gaines and Kimbrough (1968) [1] Shepelskya (1980) [1]

Fertility study

eruption and increase in the median effective time for testes descent) in the rat pups and persistent behavioral alterations (the locomotor frequency of the experimental rats was higher than that of controls at 21 and 90 days old) in adults after pregnant rats were subcutaneously treated with aldrin (1.0 mg/kg) or with its vehicle (0.9% NaCl solution plus Tween-80) from day 1 of pregnancy until delivery (Castro et al., 1992). Also aldrin may have a direct inhibitory influence on gonadotrophin release, and exert a direct action on the testes (Chatterjee et al., 1988). A review of the developmental toxicity of aldrin concludes that aldrin

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

induces malformations (eye and digit defects and cleft palate) in mice and hamsters, a low frequency of malformations in the rat; but was not teratogenic in dogs or swine. Aldrin is readily converted to dieldrin and similar results were noted for dieldrin in these species as well as in the rabbit (Schardein, 1993). In the studies submitted for registration, possible adverse effects were noted in both developmental toxicity studies (mouse and hamster at high doses) and the reproduction study in rats at the chronic toxicity dose range (CDPR, 2009). Additionally, concentrations of organochlorines such as aldrin may change over critical windows of human reproduction and development, underscoring the importance of timing biospecimen collection to critical windows for development in the assessment of reproductive and/or developmental effects (Bloom et al., 2009).

12.5.2.2  Amitraz In rats, adverse effects on reproduction were reduced litter size and substantial neonatal mortality at 200 ppm, slight to moderate neonatal mortality at 50 ppm leading to a NOAEL of 10.5 ppm. In the mouse, prolongation of the pro-estrus phase, a trend towards shortening of the diestrus phase, and depressed serum prolactin and progesterone levels with a NOAEL of 25 ppm were observed (CDPR, 2009). In a developmental neurotoxicity study, rats were administered 20 mg/kg every third day and pups born were cross fostered. Open-field behavior (locomotion and rearing frequencies or immobility time) showed no significant differences, other than some transient delays (Palermo-Neto et al., 1994). Postnatal exposure to amitraz caused transient developmental and behavioral changes in the exposed offspring in a subsequent study (Palermo-Neto et al., 1997). Results showed that the median effective time (ET50) for fur development, eye opening, testis descent, and onset of the startle response was increased in rats postnatally exposed to amitraz compared to those of the control. However, the age at incisor eruption, total unfolding of the external ears, vaginal and ear opening and the time taken to perform the grasping hindlimb reflex were not affected by amitraz exposure.

12.5.2.3  DBCP Dibromochloropropane (DBCP) is a brominated organochlorine that was used as a nematocide from the mid 1950s until its ban in the United States in the late 1970s (Whorton and Foliart, 1983). It was used in the United States mostly in Hawaii and along the southern Atlantic and Pacific coasts to protect citrus, grapes, peaches, pineapples, soybeans, and tomatoes. Of the pesticides studied to date, DBCP is the most toxic to the human male reproductive system. As early as 1961 Torkelson et al. reported

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that DBCP caused testicular atrophy in rats, guinea pigs, and rabbits. Azoospermia and oligospermia were reported in DBCP production workers and this was linked with the length of time the persons worked with the chemical (Whorton et al., 1977). Higher levels of follicle stimulating hormone (FSH) were noted in a number of these individuals as well as in those that did not revert to normal spermatogenic levels long after exposure. Data from the animal studies reveal severe testicular insults including degenerative changes in the seminiferous tubules, increase in Sertoli cells, reduction in the number of sperm, and increased abnormalities in the sperm cells. Epididymal (Kluwe, 1981), post-testicular effects (Kluwe et al., 1983), and in vitro effects (Bartoov et al., 1987) were also noted. The mechanism of action was determined to be at the level of mitochondrial respiration, and DBCP (the parent compound) was demonstrated to inhibit carbohydrate metabolism at the NADH dehydrogenase step in the mitochondrial electron transport chain of rat sperm (Greenwell et al., 1987). Workers on banana crops documented convincing evidence of increased spontaneous abortion in their family histories; follow-up studies among production workers in Israel showed that some recovered testicular function, but among their offspring, there was a predominance of females. Those who did not recover from azoospermia were those with high levels of FSH (Goldsmith, 1997). No other studies have shown increased birth defects or increased infant mortality.

12.5.2.4  DDT Reviews claim that exposure to organochlorine compounds like DDT and methoxychlor causes an impairment of female fertility by altering ovarian development and function and implantation by altering endometrial function through their estrogenic activity (Tiemann, 2008). Several investigators have examined the effects of DDT on both development and the reproductive system. DDT and its metabolite DDE have resulted in egg shell thinning in birds of different species (Porter and Wiemeyer, 1969). Thinner shells are associated with a higher disappearance and/or destruction of eggs and it is this phenomenon along with the huge public outcry subsequent to the publishing of Rachel Carson’s book Silent Spring that lead to the ban of the use of DDT in the United States. Inconsistent effects of DDT and DDE on egg shell thickness might reflect differential sensitivites among species of birds. The thinning is thought to be similar to the way estrogen inhibits formation of eggshells, although the calcium ATPase inhibition by DDT demonstrated by Matsumura and Ghiasuddin (1979) may also be responsible. The interference (direct or indirect) with fertility and reproduction is thought therefore to be related to steroid metabolism and the inability of the bird to mobilize sufficient calcium to produce a strong

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eggshell to withstand the rigors of being buffeted around the nest; and the resultant cracking allows the entry of bacteria, causing the developing embryo to die (Carson, 1962; Peakall, 1970). In male cockerels and rats, DDT (20% o,p DDT and 80% p,p DDT) reduced testicular size and in females o,p DDT administration yielded estrogenic effects such as edematous, blood engorged uteri (Ecobichon and MacKenzie, 1974; Hayes, 1959). DDT has resulted in uterotrophic effects; o,p DDT has demonstrated increased weight in ovaries and uteri and an advanced time of vaginal opening (Gellert et al., 1972). Other estrogenic effects such as increase in cellular progesterone receptors, tissue mass, DNA synthesis, and cell division have also been reported (Ireland et al., 1980; Mason and Schulte, 1980; Nelson, 1974; Robison and Stancel, 1982). Kupfer and Bulger (1976) have shown that the o,p isomer competes with estradiol for binding with estogen receptors in rat uterine cytosol. A review of studies conducted in Israel, India, and the Ukraine suggested that maternal and fetal tissue levels of DDT and metabolites were higher in fetal deaths than other pregnancies, however studies from Poland, Italy, and Florida observed no significant differences in the levels of organochlorine pesticides in samples from normal vs. aborted pregnancies (Sever, 1998). DDT is known to be lipophilic and a potent bioaccumulator and can be detected in fatty tisues in the food chain long after its use. This and the fact that it is still widely used as an efficient agent in the control of mosquitoes causing malaria can result in marked long-term ecological impact.

12.5.2.5  Methoxychlor Methoxychlor is a chlorinated hydrocarbon insecticide that has a much lower bioaccumulation potential than DDT. Early studies in pregnant rats demonstrated maternal and fetal toxicity (wavy ribs) at exposures of 200 and 400 mg/kg (Khera et al., 1978). In rats, methoxychlor is metabolized in vivo to 2,2-bis(p-hydoxyphenyl)-1,1,1-trichloroethane (HPTE), the active estrogenic form. It has direct estrogenic effects on the rat uterus and also inhibits the decidual cell response which is an accepted model for implantationassociated effects. It has adverse effects on fertility, early pregnancy, and in utero development in females; in adult males adverse effects such as altered social behavior following prenatal exposure to methoxychlor were noted (Cummings, 1997; Cummings and Gray, 1989). Recent work in mice concluded that neonatal exposure to methoxychlor at doses of 0.1, 0.5 and 1.0 mg/kg/day did not interfere with mating, but significant alterations were seen in initiating and/or maintaining pregnancy. The deleterious effects on pregnancy may be due to the influence of neonatal methoxychlor treatments on the hypothalamic–pituitary– ovarian axis as well as on possible alteration of the uterine environment (Swartz and Eroschenko, 1998). The significance of this toxicity with respect to human health

Hayes’ Handbook of Pesticide Toxicology

remains to be determined. In Long-Evans hooded rats methoxychor affects the CNS, epididymal sperm numbers, and the accessory sex glands and delays mating without significantly affecting the secretion of LH, prolactin, or testosterone. These data indicate that methoxychlor did not alter pituitary endocrine function in either an estrogenic or antiandrogenic manner (200–400 mg/ kg/day) and demonstrate a pronounced degree of target tissue selectivity (Gray et al., 1999b). In female pine voles orally administered methoxychlor throughout gestation and lactation of pups, exposed female offspring tested as adults showed a strong trend toward spending more time alone. The cingulate cortex showed a reduction in oxytocin binding demonstrating that exposure to methoxychlor during pre- and neonatal development can alter female adult neural phenotype and behavior.

12.5.2.6  Chlordecone Chlordecone was sold as an insecticide and fungicide between the years of 1958 through 1975. Better known by its tradename Kepone, it was widely used and caused contaminination of the James River near the plant where it was manufactured in Virginia. Mirex, an insecticide that photodegraded to Kepone, has also been extensively studied and the toxicity of both compounds will be summarized here. Production workers exposed to chlordecone were noted to be oligospermic and had reduced sperm motility (Taylor et al., 1978). Chlordecone primarily affects sperm motility and viability via mechanisms that are not completely understood. One study in CD-1 mice documented that the pool of potentially ovulatory follicles was reduced subsequent to prolonged exposure to Kepone (Swartz and Mall, 1989). It is a potent inducer of the mixed function oxidase system and may affect fertility by stimulating hepatic degradation of steroids. While chlordecone (probably as the hydrate) has a binding affinity for estrogenic sites (Hammond et al., 1979), other studies have concluded that the reproductive toxicity was not caused by a mimicry of estrogen (Cochran and Wiedow, 1984). Instead, it appears that chlordecone can act on the hypothalamic–pituitary axis (Hong et al., 1985). Exposure to 50 and 75 mg/kg of chlordecone in the female rat before or after mating substantially reduced fertility (Uphouse, 1986). A review of the developmental toxicity of chlordecone exposure during gestation indicated fetotoxicity in mice and rats and some central nervous system impairment in fetuses in the rat (Schardein, 1993).

12.5.2.7  Lindane (HCH) Lindane, a nonaromatic chlorinated cyclic hydrocarbon has shown some effects on the female reproductive system (Welch et al., 1971) but the results on the whole are

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

variable. In male rats fed 75 mg/kg/day of lindane for 90 days testicular atrophy, degeneration of seminiferous tubules, and disruption of spermatogenesis was reported (Shivanandappa and Krishnakumari, 1983). Reduced epididymal sperm concentrations were also noted in other studies in rats dosed a single dose of 30 mg/kg/day (Dalsenter et al., 1996). As with DDT, the estrogenic effects of lindane might occur via estrogen receptors, but studies in female Long-Evans rats dosed at 40 mg/kg/day for 7 days or in ovariectomized rats for 5 days did not show altered serum estradiol concentrations, or change in the number of estrogen receptors. No changes in estrogendependent induction of progesterone in the hypothalamus, pituitary, or uterus were observed. Hence it is thought that lindane may act via altering multiple processes such as the GABA-nergic system or via altered growth factors (Laws et al., 1994). In reproduction studies, fertility was not reduced, but most of the F1 pups died shortly after birth. Erratic estrous cycles were also noted with exposure to lindane (Gray et al., 1988). The adverse liver effects in the rat reproduction study do not appear to be important to human safety, since they were reversible after much higher exposures in the combined study (CDPR, 2009 – Record No. 091997), and the kidney effects (apparent hydronephrosis) were species and sex-specific. Reproductive NOAEL  20 ppm based on reduced neonatal pup survival (largely due to total litter losses); slightly reduced pup growth rate and slightly slowed pup development such as delays in hair growth and tooth eruption (CDPR, 2009). In a suicidal poisoning, maternal ingestion of lindane resulted in the death of twin fetuses (Konje et al., 1992). In addition to these and other endocrine effects a dose-related increase in the incidence of fetuses with an extra 14th rib in CFY rats and an extra 13th rib in rabbits has been reported at 15 mg/kg/day but a lack of teratogenicity was determined (Palmer et al., 1978). This was consistent with the negative teratogenicity results in mice, mutagenicity studies, and three-generation rat reproduction studies. Regional changes in brain norepinephrine and serotonin levels have also been reported as developmental effects (Rivera et al., 1991). Adverse effects were however not noted in the developmental toxicity studies submitted to CDPR, and currently lindane and related HCH isomers are not listed as chemicals known to the State to cause reproductive toxicity under Proposition 65 or Safe Drinking Water and Toxic Enforcement Act of 1986.

12.5.2.8  Organophosphates These compounds cause a combination of a reduction in brain acetylcholinesterase activity and altered reproductive behavior in a number of species. The reduced acetylcholinesterase has been associated with decreased egg production and serum LH and serum progesterone (Rattner et al., 1982). Commercial organophosphates and carbamates

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have demonstrated significant alteration of acetylcholine activity in vitro in human fetal brain tissue (Banerjee et al., 1991). Also experimental studies suggest that the critical period of exposure to the organophosphate pesticide chlorpyrifos extends across the pre- and postnatal periods (Qiao et al., 2002; Richardson and Chambers, 2003), with it initially acting by impairing the development of neurons, and later by affecting glia cells (astrocytes, oligodendrocytes, microglia), which are critical for brain development (Garcia et al., 2002, 2003). More recent experimental evidence suggests that subtoxic and otherwise nonsymptomatic developmental exposure to organophosphate pesticides may predispose offspring to hypertension, obesity, and diabetes (Meyer et al., 2004). The standard dominant lethal test in mice was negative for dichlorvos (Dean and Blair, 1976). Possible mechanisms of toxicity from studies on trichlorphon and parathion in the rat are thought to involve interference with steroid hormone binding to receptors in the liver, adrenal, uteri, and testes (Trajkovic et al., 1981). In a case report, the organophosphate pesticide mercarbam crossed the placental barrier and caused the death of a 5-month fetus (Tsoukali-Papadopoulou and Njau, 1987). There has been some indication that organophosphates (OPs) in general may affect the menstrual cycle and cause an early menopause in humans. Reproductive effects from exposures to mixtures of OPs have been documented by Nakazawa (1974) and Mattison et al. (1983) among women in agriculture. These effects included abnormal menstruation (e.g. hypermenorrhea, oligomenorrhea, amenorrhea), and early menopause. On the other hand, Willis et al. (1993) found no effects of pesticide exposures (including methyl parathion) on the pregnancy outcome among 535 women enrolled in a southern California community clinic perinatal program.

12.5.2.9  Chlorpyrifos The developmental and reproductive toxicity of this widely used compound have been extensively studied. The studies submitted for registration under FIFRA did not show adverse developmental or reproductive effects. However, some recent studies on the role of cholinesterase in morphogenesis have used chlorpyrifos as the model compound and intimated an influence in learning disabilities among children who were exposed in utero or during the early postnatal period (Roy et al., 1998). Chlorpyrifos, a phosphorothioate, undergoes oxidative desulfuration to form chlorpyrifos-oxon which then can phosphorylate acetyl cholinesterase rendering it incapable of metabolizing the neurotransmitter acetylcholine to choline and acetate. The oxon however may be detoxified by either combining with carboxylesterase, or may be hydrolyzed by oxonase to metabolites not capable of combining with acetylcholinesterase. It has been suggested that inhibition of DNA and protein

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synthesis may be attributed to the direct noncholinergic effects of chlorpyrifos, but other mechanisms such as alterations in blood flow patterns may also be involved. In reviewing several studies as part of the risk characterization of the compound, it appears that newborn and juvenile rodents are more susceptible to the toxic effects of chlorpyrifos than adults. The increased susceptibility of young rats appeared not to be due to a difference in the affinity of the oxon for the acetyl cholinesterase, but rather incomplete development of the enzyme systems which detoxify the oxon (CDPR, 2009). In vivo studies in rats have used newborns to try to explore the effects of chlorpyrifos on the ontogeny of the mammalian nervous system during the equivalent last trimester of human fetal development. However sublethal concentrations were administered to postnatal day 1 rats via intraperitoneal or subcutaneous routes. In the human, in utero exposures are likely to be mitigated by the mother’s metabolism and the availability of chlorpyrifos to the fetus may therefore not be comparable to newborn rodents being directly dosed. Aggregate exposure estimates, acute illness reports, and dermal exposure from surface wipes also point to low levels of chlorpyrifos (CDPR, 2009; Lewis et al., 1994; Lu and Fenske, 1999). Higher exposures to children have been reported for oral nondietary and dermal exposure (Gurunathan et al., 1998); the assumptions used, however, may not be appropriate. The inhibition of cholinesterase activity (brain and liver) in pups was detected at doses nearly lethal to the rat dams (Tang et al., 1999), hence it was thought that maternal effects will be observed prior to levels causing developmental toxicity. Nevertheless the U.S. Environmental Protection Agency (EPA) cancelled most home, lawn and garden use products containing chlorpyrifos in mid-2000 based on human health risks. In addition, as of December 2005, chlorpyrifos products were no longer permitted for use in pre-construction termite control but agricultural uses remain at the time of writing of this chapter. Epidemiological studies examining developmental outcomes associated with chlorpyrifos exposure included three prospective cohort studies, two conducted in New York City and one conducted in the Salinas Valley of California, as well as a study of agricultural exposure in India. Adverse outcomes reported in association with exposure to chlorpyrifos included: lower birth weight; decreased birth length; decreased head circumference, association with paraoxonase (PON1) status; mental and motor delays; behavioral problems; as well as DNA damage. Study parameters differed and may account for some of the inconsistencies in the findings. Of the several studies in laboratory animals some standard developmental toxicity studies have reported effects on survival, growth and maturation, although these findings are generally equivocal. In other studies that include endpoints not covered by standard developmental toxicity studies, neurochemical alterations after birth and resultant behavioral effects have

Hayes’ Handbook of Pesticide Toxicology

been reported to occur after exposure during the gestational period. The underlying mechanism of neurotoxicity is not fully understood. Effects reported in a developmental neurotoxicity study include alterations in motor activity, auditory startle response, and brain structure (decreased measurements of the parietal cortex and hippocampal gyrus in the absence of significant brain weight deficits at 5 mg/ kg/day). After exposure to 1 mg/kg/day, female offspring also exhibited significant dose- and treatment-related decreases in measurements of the parietal cortex at postnatal day (PND) 66, long after exposure to chlorpyrifos had ended. This indicates an association of chlorpyrifos with delayed alterations in brain development in offspring of exposed mothers who showed minimal plasma and erythrocyte cholinesterase inhibition. Oral exposure of dams to 3 mg/kg/day chlorpyrifos was the lowest dosage that resulted in significant (10%) inhibition of brain cholinesterase in the offspring on postnatal day 1. In some studies, prenatal exposures that were nontoxic to the dam elicited deficits in cholinergic function in offspring that influence cognitive performance in adolescence and adulthood. In other studies, prenatal chlorpyrifos exposure appears to elicit delayed-onset alterations, disrupting the program for the emergence of cholinergic activity. Thus, although the studies raise many questions regarding possible mode of action, dosing and behavioral assessments with different behavioral techniques, when taken together they provide a basis for concern for susceptibility for persistent effects of chlorpyrifos on neurodevelopment.

12.5.2.10  Diazinon This organophosphate insecticide has been tested extensively and yielded variable results for reproductive and developmental endpoints. Spyker and Avery in 1977 exposed pregnant mice (9 mg/kg) and observed behavioral effects and functional impairments in overtly normal offspring along with neuropathology in the forebrain. The standard developmental toxicity studies in rats and rabbits do not demonstrate adverse effects (CDPR, 2009). In another review malormations were not reported in hamsters and rabbits, but renal, rib, limb and anomalies of the central nervous system and digits were noted in rats, and skull and teeth abnormalities were noted in puppies (Schardein, 1993). In both one-generation and two-generation reproduction studies, a variety of adverse effects were observed. These include a decrease in the gestation index (number of litters with live offspring/number pregnant), a reduction in ovarian weights, and a prolonged gestation length. The reproduction NOAEL 10 ppm (1 mg/ kg/day) or LOEL  10 ppm (CDPR, 2009). Behavioral effects were confirmed in neurotoxicity studies in rats (CDPR, 2009). Recently in rats, neonatal diazinon exposure, at doses below the threshold for cholinesterase inhibition, demonstrated effects on emotional responses altering serotonergic synaptic function. The effects were preferential in

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

males with no significant effects for females. The outcomes examined reflect different aspects of emotionality such as: (1) decreased venturing into the open arms in the elevated plus maze – typically interpreted as increase in anxiety; (2) loss of preference for sweetened liquids – characteristic of anhedonia, a typical component of animal models of depression (Roegge et al., 2008). Some inhibition of acetyl cholinesterase with a higher dose reversed the cognitive impairment and this nonmonotonic dose–effect function has also been seen with neurochemical effects. While some of the diazinon effects on cognition resemble those seen earlier for chlorpyrfos, some differ suggesting that diazinon and chlorpyrifos affect transmitter systems supporting memory function differently.

12.5.2.11  Dimethoate Developmentally toxic effects were not noted in either the rat or rabbit studies submitted under FIFRA (CDPR, 2009), however rib defects in rats and polydactyly in cats have been noted (Khera, 1979). In the mouse, variable results are noted with embryotoxicity without teratogenicity in earlier studies (Schuefler, 1975, 1976) and no adverse effects in another study (Courtney et al., 1985). While an absence of anomalies in fetuses (gross, visceral morphology and skeleton) was noted in rats (Srivastava and Raizada, 1996), reduced number of pregnant females and lowered pup weights and reduced litter sizes have been noted in a two-generation reproduction study in rats (CDPR, 2009). More recent studies have demonstrated a decrease in thyroid hormones (free T4 and T3) in suckling rats exposed to dimethoate (Mahjoubi-Samet et al., 2005).

12.5.2.12  Fenthion A reduction in fetal weights at 80 mg/kg was noted in a study in mice with an increase in malformations in 14.5% of the offspring (Budreau and Singh, 1973a). In rats, however, a marginal increase in resorptions was noted at 18 mg/kg/ day demonstrating no other adverse developmental effects at lower doses, yielding a NOAEL  4.2 mg/kg/day. Exposure to fenthion in the diet at 14 and 100 ppm in the reproduction study demonstrated epididymal cytoplasmic vacuolation (ECV) associated with decreased fertility, reduced survivability, and postnatal growth retardation resulting in a reproductive NOAEL  2 ppm (CDPR, 2009).

12.5.2.13  Parathion and Methyl Parathion Three multigeneration reproductive toxicity studies in rats have been submitted to regulatory agencies (CDPR, 2009; RCD) and decreased pup survival was consistently found in all three studies. A search of the literature revealed one study showing ovarian toxicity in rats (Dhondup and Kaliwal, 1997) and one study showing possible sperm abnormalities in mice (Mathew et al., 1992). Testicular

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and reproductive effects have also been reported in avian species (Maitra and Sarkar, 1996; Solecki et al., 1996). In the absence of clinical symptoms and behavioral changes, reproductive effects such as reduction in the number of eggs laid (~20% reduction), egg weight (~9% reduction), and eggshell thickness (7–10% reduction) were noted in the Japanese quail. Suppression of growth and ossification in both mice and rats were observed subsequent to methyl parathion exposure; in the mouse high mortality and cleft palate were noted (Tanimura et al., 1967). No other epidemiological data specific to methyl parathion are available.

12.5.2.14  Malathion Malathion does not appear to cause adverse developmental or reproductive effects (CDPR, 2009). However, malathion administered to mice at 250 mg/kg (corresponding to 121 LD50) and examined 4, 14, 18, and 26 days after injection induced teratozoospermia. Sperm count at different time intervals was significantly increased compared to controls and there was a parallel increase in depletion of the seminiferous tubules; all germinal cell populations studied were affected by malathion, especially mice spermatid differentiation, which compromises mostly the flagella, perhaps due to an alkylating effect that disturbs the normal assembling of tail structural protein components (Contreras and Bustos-Obregon, 1999). No evidence for the histopathological alteration or teratogenic anomalies in the fetuses were demonstrated, though placental transfer of malathion was indicated by the presence of the insecticide residues in fetuses from rats fed wheat material containing bound residues of malathion S-1,2-di(ethoxycarbonyl) ethyl O,O-dimethyl phosphorodithioate (Bitsi et al., 1994). The reproductive effects of the aerial spraying of the organophosphate insecticide malathion in California have been examined in a case-control study of spontaneous abortions (28 weeks) and stillbirths; relative risks were 1.21 (95% CI  0.94–1.52) for spontaneous abortions and 1.51 (95% CI  0.21–11.3) for stillbirths. A cohort of 7450 pregnancies identified through three Kaiser-Permanente facilities in the San Francisco Bay Area, in relation to exposure to the pesticide malathion, applied aerially to control an infestation by the Mediterranean fruit fly was examined for reproductive outcomes. No important association was found between malathion exposure and spontaneous abortion, intrauterine growth retardation, stillbirth, or most categories of congenital anomalies. Gastrointestinal anomalies noted were related to second trimester exposure (OR  2.6), based on 13 cases and not specific to any particular International Classification of Diseases code (Thomas et al., 1992). Salazar-Garcia et al. (2004), in a more recent study evaluating the reproductive history of workers in an antimalaria campaign in Mexico, noted that exposure to malathion increased the risk of birth defects after controlling for DDT exposure (OR  2.06; 95% CI  1.01–4.22).

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12.5.2.15  Trichlorfon Trichlorfon, an organophosphate insecticide, has been associated with a cluster of babies born with Down syndrome in Hungary (Czeizel et al., 1993). A case-control study and environmental investigations reported excessive use of trichlorfon at local fish farms. The high content of trichlorfon in the diet of pregnant women including all of the mothers with affected offspring, along with absence of known teratogenic factors such as familial inheritance and consanguinity, was supportive of the association. Trichlorfon or chlorophos, marketed under the brand name Dipterex, has embryotoxic and teratogenic effects in the Wistar rat after an oral dose of 80 mg/kg during a critical period of embryogenesis, but was negative at the low dose of 8 mg/kg (Martson and Voronina, 1976). A review of this insecticide and anthelmintic has demonstrated a potent developmental toxicity profile in laboratory animals such as the rat (oral and inhalation), mouse, and hamster (oral), but teratogenicity was not noted if the exposure was via the intraperitoneal route (Schardein, 1993). Recent work from Norway has characterized teratogenic effects of trichlorfon on the guinea-pig brain, by determining the effective dose and sensitive period (Hjelde et al., 1998). Following oral or subcutaneous exposure, almost all regions of the brain were reduced in weight. The cerebellum was the most vulnerable region, but the medulla and hypothalamus were also greatly reduced in weight. While the mechanism behind the teratogenic effect is not known, alkylation of DNA or altered DNA repair may be involved.

12.5.2.16  Carbamates Similar to organophosphates, carbamates result in the inhibition of cholinesterase and also exert an anesthetic effect. The dithiocarbamates have been used as fungicides and will be discussed under that category.

12.5.2.17  Carbofuran Most studies with carbofuran have been negative for teratogenicity. However one study in mice resulted in fine structure abnormalities in mice (Schardein, 1993) and in the FIFRA reproduction study, reduced body weight gain in adults and birthweights in offspring worsening to 15% by weaning were noted (CDPR, 2009). Decreased weights of the epididymides, seminal vesicles, ventral prostate, and coagulating glands were also noted in rats, along with decreased sperm motility, reduced epididymal sperm count and increased morphological abnormalities in head, neck and tail regions of spermatozoa (Pant et al., 1995). Testicular and spermatotoxic effects were also noted at levels higher than 0.2 mg/kg in rats exposed to carbofuran in utero or via lactation (Pant et al., 1997). Studies from Sri Lanka in rats have concluded that carbofuran administered

Hayes’ Handbook of Pesticide Toxicology

orally at 0.2, 0.4 and 0.8 mg/kg during early gestation is detrimental to pregnancy (enhanced preimplantation losses) and possibly harmful to neonatal development (Jayatunga et al., 1998a). Similarly postimplantation losses were noted after exposure to carbofuran during mid-gestation (Jayatunga et al., 1998b).

12.5.2.18  Bendiocarb This compound is a residual insecticide and appears to cause adverse effects on reproduction in the rat. A decrease in number of pups and reduced survival at 200 ppm and 250 ppm dose levels resulted in a NOAEL of 50 ppm (CDPR, 2009).

12.5.2.19  Thiodicarb Adverse effects were documented in the rat reproduction study submitted for registration. Decreased pup weight gain at 100 ppm and above and decreased viability index at 900 ppm was noted. Parental NOAEL  100 ppm (decreased F0 and F1 body weights). Reproductive LOAEL  100 ppm (reduced pup weight gain). A statistically based estimate of a NOAEL provided a NEAEL (No-Expected-Adverse Effect-Level)  81 ppm in males; 80 ppm in females (CDPR, 2009).

12.5.2.20  Zineb and Thiram The exposure of rats to zineb and thiram has documented an alteration (prolongation) of the estrous cycle in association with a reduction in ovarian and uterine weights (Ghizelea and Czeranschi, 1973). Direct gonadal effects have been noted in mouse and rabbit oocytes resulting in inhibition of oocyte meiotic maturation and prevention of germinal vesicle breakdown. In the mouse oocyte exposed to isopropyl-N-phenylcarbamate, a formation of nuclear condensates (macromolecules) has also been observed by Crozet and Szollosi (1979). Decreased body weight in dams and pups was noted during gestation and through lactation at 180 ppm in the F0 and at 60 ppm and above in the F1 generation. Hence developmental effects were noted resulting in a NOAEL of 20 ppm in a reproduction study in rats exposed to thiram in the diet (CDPR, 2009).

12.5.2.21  Carbaryl Carbaryl (N-methyl-1-naphthyl carbamate; Sevin) is a carbamate insecticide that is widely used, resulting in exposure during its use as well as via consumption of treated food. It is not metabolized to an active intermediate, the parent compound itself is thought to be the active agent. Carbaryl acts via inhibition of acetylcholinesterase by carbamylation of the active-site serine residue. Adverse events on rodent spermatogenesis at 0.4–5 mg/kg i.p. or p.o. have been reported (Weil et al., 1972), but several studies do

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

not support this finding. Hence, although the data suggest that carbaryl does not produce a testicular effect similar to DBCP, human personnel may be affected as demonstrated by some studies (Wyrobek et al., 1981). The teratogenicity of carbaryl has been reviewed extensively but the results can best be summarized as equivocal (Schardein, 1993). Data are available in a number of species; early studies in the rat reported terata but subsequent studies were negative. Studies in mice were variable and the eye defects seen previously were not observed in later studies. In rabbits the data are contradictory, omphalocele and skeletal variations were noted at the high dose, however malformations were not seen in the other study. A recent study in rabbits submitted to CDPR documented agenesis of the gall bladder at the high dose in two fetuses of different litters along with reduction in size of the gallbladder and a missing bile duct presenting the case for a continuum of effects (CDPR, 2009). Sheep that were fed carbaryl in the diet demonstrated heart defects and in the dog where pregnant females were dosed with 6.25–50 mg/kg/day in the diet defects included abdominal–thoracic fissures, intestinal agenesis and displacement, brachygnathia, failure of skeletal formation, anurous (no tail), and superfluous phalanges. In addition to these malformations that were seen in several pups, resorption was noted in 21 of 181 pups (11.6% fetal incidence; 21.1% litter incidence). However per the earlier mentioned review, exposure to carbaryl during gestation resulted in no malformations in primates, minature swine, and cattle, though abortions were noted in the primate study. In a dietary developmental neurotoxicity study with carbaryl as reported by Sette (2001), alterations in several internal brain measurements were seen in postnatal day 11 pups at the highest dose tested (10 mg/kg/day), but pup brains at the mid and low dose were not examined.

12.5.2.22  Pyrethroids Pyrethroid pesticides are the synthetic analogs of the naturally occurring toxin, pyrethrin, derived from the flowers of Chrysanthemum cinerariaefolium. These are considered relatively safe and are perceived to be innocuous because of the origin of the natural pyrethrins from the chrysanthemum family of plants. Pyrethroids exert their toxic action by binding to the voltage-dependent sodium channel in nervous tissue and prolonging the open phase (Soderlund et al., 2002; Vijverberg and van den Bercken, 1990). While these pesticides have been modified to be more photostable, lipophilic, and more toxic than pyrethrin, they are considerably less toxic to mammals than other classes of insecticides and are widely used to control insects in and around homes and child care facilities. Several pyrethroids have been detected in floor wipe samples taken from 168 daycare centers, as residues in selected baby food samples, in human breast milk, and in urine samples of children (Pine et al., 2008). Earlier studies on the metabolism and

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toxicity of synthetic pyrethroids (fenothrin, furamethrin, proparthrin, resmethrin, tetramethrin, and allemethrin) indicate that neither the cis nor trans isomers of chrys­ anthemumate is teratogenic in rats, mice, and/or rabbits (Miyamoto, 1976). Toxicity studies with decamethrin, a synthetic pyrethroid, found no evidence of teratogenic activity in rats or mice at dose levels that produced marked maternal toxicity (Kavlock et al., 1979). However, numerous studies on the genetic toxicity potential of this group of compounds (cypermethrin and deltamethrin) have demonstrated a wide range of effects including mitotic/chromosomal abnormalities and the induction of sister chromatid exchanges (Chauhan et al., 1997). In the studies submitted in support of registration for deltamethrin (as noted in the Risk Characterization Document), no significant developmental toxicity was reported in rats; delayed ossification was noted in the high dose in rabbits along with maternal effects (CDPR, 2009). However a 5% deltamethrin formulation in Wistar derived albino rats resulted in dose-dependent early embryonic death, retardation of fetal growth, hypoplasia of the lungs and dilation of the renal pelvis with no skeletal abnormalities (Abd El-Khalik et al., 1993). Significant increases over respective controls were evident for chromosome aberrations, micronuclei, or sperm abnormalities (Bhunya and Pati, 1990). Several in vitro studies have indicated that pyrethroids may have estrogenic activity, causing them to be placed on the Environmental Protection Agency’s list of possible endocrine disruptors, and fenvalerate has been shown to induce proliferation and increase the expression of the estradiol-inducible gene, pS2. Additionally while no evidence of additional sensitivity to young rats or rabbits following pre- or postnatal exposure to esfenvalerate was determined by regulatory agencies, studies show that immature female rats exposed to 1.0 mg/kg/day are sensitive as evidenced by their delay in the onset of puberty (Pine et al., 2008). In addition to a delay in sexual maturation (while body weight was unchanged), abnormal estrous cycles and a reduction in sexual behavior were noted subsequent to maternal exposure to fenvalerate (FV) during the prenatal and postnatal periods of sexual brain organization of female offspring (Moniz et al., 2005). However gonadal hormone levels in the plasma, stereotypy and open-field behaviors were not affected. A number of effects of exposure to pyrethroids during early development have been described in rats and mice. Cypermethrin caused an apparent increase in blood–brain barrier permeability in 10-day-old rat pups after a single dose or repeated doses of about 15% of the LD50 but had no effect on the adult barrier (Gupta et al., 1999). A slightly lower daily dose of 4% of the LD50 over postnatal days 10–16 caused an increase in renal D1 receptor density in rats which persisted at least until day 90 (Cantalamessa et al., 1998). Similarly low doses of bioallethrin to mice over postnatal days 10–16 decreased muscarinic receptor

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density in adult mouse neocortex and produced lasting changes in adult behavior (Talts et al., 1998). Subsequent work by these researchers suggest that exposure to specific pesticides including bioallethrin during the brain growth spurt period in mice can potentiate susceptibility to bioallethrin or paraoxon in adult life (Eriksson and Talts, 2000). The pyrethroids are also capable of producing gross effects on brain maturation and morphology, but only if given at dose levels which cause reduced body weight in the offspring (Patro et al., 1997), probably via a nonspecific developmental delay due to undernutrition. Additionally prenatal exposure to low doses (0.25 or 0.5 or 1.0 mg/kg, p.o.) of deltamethrin, a type II pyrethroid insecticide, to pregnant dams from gestation days 5 to 21 (GD5–21) produced dose-dependent alterations in the ontogeny of xenobiotic metabolizing cytochrome P450 (CYP) isoforms in brain and liver of the offspring (Johri et al., 2006). Since these CYP enzymes have a role in regulating the levels of ligands that modulate growth, differentiation, and neuroendocrine function, these findings are important. Findings from exposure to pyrethroid-based mosquito repellent inhalation during the early developmental period suggest adverse effects on the developing nervous system causing cholinergic dysfunction leading to learning and memory deficit (Sinha et al., 2006).

12.5.3  Insect Growth Regulators 12.5.3.1  Methoprene A review of the developmental effects of methoprene indicated that a high incidence of multiple malformations was induced in mice, but not rats (Schardein, 1993). In 1995, middle school students reported (on the internet) a high incidence of malformed frogs from a southern Minnesota farm pond. Consequently, increased rates of congenital anomalies in regions in Minnesota associated with pesticide use has heightened the attention to the possible effect of pesticides (Garry et al., 1996). Another group has implicated agricultural contaminants in the hindlimb deformities in frogs from a number of ponds in Quebec (Ouellet et al., 1997). Additionally some degradation products of the insect growth regulator S-methoprene have been reported to alter early frog embryo development in the laboratory (La Clair et al., 1998). However, confirmation of these effects in mammlian species is lacking. The standard teratogenic studies conducted under FIFRA requirements for methoprene do not demonstrate similar results. Further, it is not known if this compound, a juvenile growth hormone agonist, is used in quantities high enough to be a cause for concern. Recent findings linking the limb defects in frogs to a trematode parasite has shifted the suspicion from methoprene (Ankley et al., 1998), but in the interest of providing the reader with the putative effects of this compound the above-mentioned information is provided.

Hayes’ Handbook of Pesticide Toxicology

12.5.3.2  Diflubenzuron Other insect growth regulators such as diflubenzuron (Dimilin, TH 6040; N-[[(4-chlorophenyl)amino]-carbonyl]-2, 6-difluorobenzamide) have been tested in male and female layer-breed chickens from 1 day of age through a laying cycle at levels of 1, 2.5, 25, and 250 ppm in the feed. Feeding diflubenzuron at levels up to 250 ppm did not affect the characteristics measured such as egg production, egg weight, eggshell weight, fertility, hatchability, and effects on the progeny (Kubena, 1982).

12.5.3.3  Fenoxycarb Fenoxycarb (ethyl [2-(4-phenoxyphenoxy)-ethyl] carbamate) is an insect growth regulator used for long-term fire ant control. In a reproductive study Rambouillet sheep were dosed daily with a placebo or with fenoxycarb at 0.69 or 1.38 mg/kg/day, representing 10 (10) and 20 times (20) the maximum amounts of fenoxycarb in forage or hay treated at recommended levels for fire ant control. No statistically significant (p  0.05) differences were seen between the exposed and control groups of sheep for rates of weight gain of adults, serum clinical chemistry profiles of adults, spermatozoa morphology and motility, estrus cycling, pregnancy rates, maintenance of pregnancies to term, numbers of livebirths, and rates of weight gain of lambs to 28 days. No clinical signs associated with exposure to fenoxycarb were observed in any animal at any time, and no exposure-related pattern of pathologic lesions or reproductive organ histology was observed. Based on the lack of significant findings in this study, it is unlikely that use of fenoxycarb, according to label instructions (currently applicable to homeowner and registered agricultural usage) for fire ant control in pasturage or hay fields will affect ruminant reproduction (Barr et al., 1997).

12.5.4  Fungicides (Table 12.5) Most fungicides tend to produce positive results in the standard in vitro microbial mutagenicity tests. This is because the microorganisms used in such test systems are similar to the fungi. However given the predictive possibility of the mutagenicity tests for teratogenic and carcinogenic potential, there is mounting concern about the developmental toxicity of these compounds. Several fungicides have documented developmental toxicity and details are available below. There is evidence to suggest that fluconazole, a bis-triazole antifungal agent, exhibits dose-dependent teratogenic effects; however, it appears to be safe at lower doses (150 mg/day). Ketoconazole, flucytosine, and griseofulvin have been shown to be teratogenic and/or embryotoxic in animals. Iodides have been associated with congenital goiter and should not be used during pregnancy.

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

415

Table 12.5  Developmental and Reproductive Toxicity Profile of Fungicides Dose (mg/kg)a

Chemical

Species

Toxicity profile

Alkyldithiocarbamic acid

Rat

Teratogenic

Benomyl

Mouse

Developmentally toxic Teratogenic

100

Rat

Fetal mortality Teratogenic Reproduction

200 200

Rabbit

Visceral variations

180

Munley and Hurtt (1996)

Mouse

Developmentally toxic Teratogenic

23.4

Davis et al. (1987)

Rat

Developmentally toxic Teratogenic

10 10

Bitertanol

Rat

Teratogenic

Captafol

Rat

–

500

Rabbit

–

150

Primate

–

25

Hamster

Teratogenic

200

Vondruska et al. (1971) [1]

Mouse

–

100

Robens (1970a) [1]

Rat

–

2000

Courtney et al. (1978) [1]

Rabbit

Teratogenic

37.5

Kennedy et al. (1972) [1]

Primate

–

75

Hamster

Teratogenic

300

Dog

Teratogenic

Rat

Embryotoxic Teratogenic

100 200

Robens (1974)

Rabbit

–

160

Cummings et al. (1992)

Rat

–

1.5 g/100 g BW

s.c. route

Janardhan et al. (1984)

Rabbit

–

47 g/animal

i.v. route

Marois and Buvet (1972) [1]

Rat

Developmentally toxic Teratogenic Toxic to testis

0.01

Mouse

Teratogenic

30

Schardein (2000)

Rabbit

Developmentally toxic

0.05

Lary and Hood (1978) [1]

Rat

–

400

Schardein (2000)

Bis(tri-N-butyltin) oxide

Captan

Carbendazim

Cupric acetate

Cycloheximide

1 20

Comments

References Petrova-Vergieva (1971) [1]

LD50

Kavlock et al. (1982) Selectively toxic Teratogenic p.o.; negative via diet 3-G study

Ruzicska et al. (1975) [1] Kavlock et al. (1982) Sherman et al. (1975) [1]

11.7

1 10

Crofton et al. (1989)

Vergieva (1990)

LD50

Kennedy et al. (1972) [1]

2 species

McLaughlin et al. (1969) Vondruska et al. (1971) [1] Robens (1970a) [1]

0.01

(Continued )

Hayes’ Handbook of Pesticide Toxicology

416

Table 12.5  (Continued) Dose (mg/kg)a

Chemical

Species

Toxicity profile

Cymoxanil

Rabbit

Teratogenic, early resorptions, skeletal variations

Varnagy and Imre (1980) [1]

Dazomet

Rat

–

CDPR Toxicology Summary (1998) DPN # 50466

Dinocap

Mouse

Developmentally toxic

12

Selectively toxic

Gray et al. (1986)

Rat

Developmentally toxic Reproduction

100

By both oral and dermal routes 4-G study

Rogers et al. (1988) Fraczek (1979) [1]

Rabbit

Fetotoxic

48

By oral and dermal routes

Costlow et al. (1986)

Hamster

Developmentally toxic

12.5

Rogers et al. (1988)

Ethylenebisisothiocyanate sulfide

Mouse

–

200

Chernoff et al. (1979a)

Rat

Functional alterations

30

Ethylthiuram monosulfide

Rat

–

60

Ruddick et al. (1976)

Ferbam

Mouse

–

300

Minor et al. (1974) [1]

Rat

Teratogenic Reproduction

150

Minor et al. (1974) [1] Vettorazzi (1975) [1]

Flusilazole

Rat

Teratogenic

1/5 LD50

Vergieva (1990)

Folpet

Mouse

–

100

Courtney et al. (1978) [1]

Rat

–

500

Kennedy et al. (1972) [1]

Rabbit

–

80

Fabro et al. (1966) [1]

Hamster

Teratogenic

500

Robens (1970a) [1]

Primate

–

2 species

Vondruska et al. (1971) [1]

Mouse

Teratogenic

100

Courtney et al. (1976)

Rat

Developmentally toxic Developmentally neurotoxic Reproduction

10

Khera (1974) [1]

Rabbit

–

10

Villeneuve et al. (1974) [1]

Mouse

Teratogenic, increased skeletal defects

10

CDPR Toxicology Summary (1995) DPN # 413

Rat

Teratogenic, resorptions, reduced fetal weights

NOAEL  40

Reproduction: decreased litters and litter size

Pup NOAEL  20

Hexachlorobenzene

Imazalil

Comments

References

Goldey and Taylor (1992) Grant et al. (1977) [1]

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

417

Table 12.5  (Continued) Chemical

Species

Toxicity profile

Dose (mg/kg)a

Imidazolidinethione

Mouse

–

800

Teramoto et al. (1978)

Rat

Developmentally toxic Teratogenic

10

Chernoff et al. (1979b)

Rabbit

Resorption

10

Khera (1973c)

Hamster

Teratogenic

270

Teramoto et al. (1978)

Cat

Teratogenic

5

Khera and Iverson (1978)

Isoprothiolane

Mouse

Developmentally toxic

100

Sakurai and Kasai (1976) [1]

Mancozeb

Mouse

–

1330

Larsson et al. (1976)

Rat

Teratogenic

1330

Rabbit

–

80

Solomon and Luiz (1989)

Mouse

Increase in variations Altered behavior

375

Chernoff et al. (1979b) Morato et al. (1989)

Rat

Teratogenic Reproduction

480

Chernoff et al. (1979b) Vettorazzi (1975) [1]

Metiram

Rat

Teratogenic, decreased live litter size

80

CDPR Toxicology Summary (1991) DPN # 217

Ochthilinone

Rat

–

56

Costlow et al. (1983)

Rabbit

Embryotoxic

1.5

Phenoxyacetic acid

Mouse

–

900

Hood et al. (1979a) [1]

Phenylphenol

Mouse

–

2100

Ogata et al. (1968) [1]

Rat

Developmentally toxic Reproduction

150

Teramoto et al. (1977)

Rabbit

–

250

Zablotny et al. (1992)

Polycarbacin

Rat

Embryotoxic Teratogenic

610

Martson and Martson (1970) [1]

Propamocarb

Rat

Teratogenic, skeletal changes

0.31 ml/kg

CDPR Toxicology Summary (1998) DPN # 50308

Propineb

Rat

Teratogenic

1000

Petrova-Vergieva (1976) [1]

Sodium phenylphenol

Mouse

–

400

Ogata et al. (1978) [1]

Terrazole

Rabbit

Increased resorptions, malformations

15

CDPR, Toxicology Summary (1990)

2,3,4,6-Tetra chlorophenol

Rat

Delayed ossification

30

Schwetz and Gehring (1973) [1]

Thiophanate ethyl

Mouse

Retarded growth Reproduction

200

Maneb

20

Comments

References

Selectively toxic

Eigenberg et al. (1997) [2]

Makita et al. (1970b) [1] 3-G study (Continued )

Hayes’ Handbook of Pesticide Toxicology

418

Table 12.5  (Continued) Chemical

Species

Toxicity profile

Dose (mg/kg)a

Thiram

Mouse

Resorption Teratogenic

10 mg 10 mg

Matthiaschk (1973) [1]

Rabbit

Mortality Teratogenic

0.01 LD50 0.01 LD50

Zhavororkov (1979) [1]

Hamster

Teratogenic

250

Robens (1969) [1]

2,4,5-Trichlorophenol

Mouse

–

900

Hood et al. (1979a) [1]

Tridemorph

Mouse

Teratogenic

245

Merkle et al. (1984)

Rat

Developmentally toxic Teratogenic

60.2 60.2

Selectively toxic Typical of endocrine disruption

Comments

References

Vinclozolin

Rat

Reproductive malformations

100

Zineb

Rat

Teratogenic

2 g/kg

Petrova-Vergieva and Ivanova-Chemishanska (1973) [1]

Ziram

Rat

–

250

Nakaura et al. (1984)

Gray et al. (1994)

Notes: Many of the citations provided in the References column have been taken from the textbook Chemically Induced Birth Defects by Schardein (1993, 2000). These references are indicated by either [1] or [2] following the citation, for Schardein (1993), and Schardein (2000), respectively. All CDPR citations may be accessed at the departmental webpage at http://www.cdpr.ca.gov/docs/risk/toxsums/toxsumlist.htm. ‘Developmentally toxic’ includes reduced fetal weight, increased embryo/fetal mortality and/or increased developmental variations, but does not include malformations. Teratogenic is the term used to indicate presence of malformations. a Doses are oral unless stated otherwise; is LOAEL wherever effects were observed and NOAEL when there were no effects.

12.5.4.1  Benomyl In reproduction studies in the rat a reduction in epididymal sperm counts in pubertal animals was observed. Postpubertal animals showed a wide variation in susceptibility of sperm counts. Histological exams of testicular tissue showed an increased incidence of diffuse hypospermatocytogenesis in pubertal and postpubertal males (Carter et al., 1984). In the 3000 and 10,000 ppm males, lower sperm counts were noted. Also testicular atrophy and degeneration (4/30 and 29/30 in P1 and 9/30 and 21/25 in F1 3000 and 10,000 ppm groups respectively), and oligospermia in the epididymides (unilateral and bilateral with 1/30 at 3000 ppm and 26/30 at 10,000 ppm in P1, 9/30 and 20/25 in F1 respectively) were observed. For the reproduction study, the NOAEL  500 ppm in males and 3000 ppm in females (decreased body weights). The NOAEL for developmental toxicity  31.2 mg/kg/day (dose-related reduction in fetal weight, hydrocephaly, microphthalmia, fused ribs, fused vertebrae, and decreased ossification in tail and in vertebral centra) in rats. Findings at the highest dose tested of 125 mg/kg/day included: full litter resorptions in 6 of 11 surviving pregnant dams, enlarged lateral ventricles, enlarged renal pelves, and delayed ossification (more widespread than at 62.5 mg/kg/day). Fetotoxicity and

teratogenicity findings in the absence of obvious maternal toxicity indicate possible adverse effects. MBC, a metabolite of benomyl, appears to cause significant effects (postimplantation loss) in rabbits at the mid and high dose level and resulted in a developmental NOAEL  10 mg/kg/day vs. maternal NOAEL  20 mg/kg/day. However, a teratology study in rabbits exposed to benomyl that was acceptable to CDPR did not demonstrate possible adverse effects (CDPR, 2009).

12.5.4.2  Dinocap Technical-grade dinocap, a complex-mixture fungicide, has been demonstrated to be teratogenic in the CD-1 mouse, causing cleft palate, a dose-related increase in supernumerary ribs, a low frequency of exencephaly, umbilical hernia at high doses, otolith defects, weight deficits in fetuses at term, increased neonatal mortality, abnormal swimming behavior, and torticollis (Rogers et al., 1986). Neither of the purified isomers 2,4-dinitro-6-(1-methylheptyl)phenyl crotonate and 2,6-dinitro-4-(1-methylheptyl)phenyl crotonate, exhibited any developmental toxicity when administered under identical conditions (Rogers et al., 1987). Similar developmental defects were not seen in the rat and hamster (Rogers et al., 1988).

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

12.5.4.3  Folpet Several studies on folpet have shown variable results relative to potential developmental effects. Although the differences in results between these studies may be due to variation in individual susceptibility, the weight of evidence indicates that in the rat, a possible adverse developmental effect (reduced ossification of interparietal bones and incidence of angulated ribs) is present. An estimated developmental NOAEL for calculation of margins of safety is 15 mg/kg, one-tenth of the lowest observed effect level (CDPR, 1989).

12.5.4.4  Hexachlorobenzene Hexachlorobenzene (HCB) is a pre-emergent fungicide and is ubiquitious in the environment. It has been isolated in the repoductive tract in several species including humans (Jarrell et al., 1998; Trapp et al., 1984). Although HCB was not mutagenic in microbial test systems and was negative in dominant lethal mutation tests, it did cause terata in mice (renal and palate malformations) and in rats (increased incidence of the 14th rib). HCB was also found to be particularly toxic to the developing perinatal animal, transplacentally and via the milk causing enlarged kidneys, hydronephrosis, hepatomegaly, and possible effects on the immune system (Ecobichon, 1996). It is commonly present in fat because of its lipophilicity and tendency to bioaccumulate (Mes et al., 1982). Adverse reproductive effects of HCB have been reported in rats, minks, ferrets, and monkeys (Bleavins et al., 1984; Iatropoulos et al., 1976). These effects include decreased fertility, fecundity, and impaired cyclicity. In a recent study in Germany, HCB concentration correlated with maternal age (r  0.249; p  0.01), with 2.7-fold higher serum levels in offspring of 40-yearold as compared with 20-year-old women, concluding that the neonatal burden depends on maternal age and duration of pregnancy. This reflected the increase in body accumulation with these substances during human life as well as a continuous transplacental transfer from mother to fetus during pregnancy (Lackmann et al., 1999).

12.5.4.5  Ethylene Dibromide (EDB) 1,2-Dibromoethane, primarily a scavenger of lead compounds in gasoline, has also been used extensively as a fumigant for its chemical and biocidal properties as a soil sterilant and a spot fumigant or control agent in grain milling machinery, grain, and in fruit and vegetable infestations. In addition to its tumor-causing capabilities in rats and mice; it has been documented to cause changes in sperm morphology in bulls (Amir and Volcani, 1965). Spermatids appear to be the target for this compound and it has been shown to affect spermatogenesis in rat, bulls and rams and to affect fertility in fowl (Alexeeff et al., 1990). Human studies indicate that EDB may harm sperm and

419

decrease fertility. While it is a reproductive toxin, it does not appear to be teratogenic.

12.5.4.6  Ketoconazole Ketoconazole, an imidazole antifungal agent, can compromise early pregnancy and also affects P450 enzymes of the mammalian steroidogenic system and inhibits progesterone synthesis in the ovary (Cummings et al., 1997). It is a potential anti-androgenic agent and has displayed anti-hormonal activities, apparently by inhibiting ovarian hormone synthesis, resulting in delayed delivery and whole litter loss (Gray et al., 1999a).

12.5.4.7  Triazoles Triazole fungicides such as propioconazole and imazalil may affect pregnancy by inhibiting aromatase (CYP 19), an enzyme critical for successful pregnancy maintenance and one that converts androgens to estrogens. Recent findings by Sanderson et al. (2002) have shown that imidazole fungicides, such as imazalil, are potent inhibitors (IC50 0.1 M) of aromatase in H295R cells (from a human adrenocortical carcinoma cell line) and appeared to act through a mixed competitive/noncompetitive mechanism. Triazole fungicides, such as propiconazole, were also inhibitors (IC50 5 M) of aromatase activity, acting competitively. The mechanism of induction appeared to involve an increased level of cyclic AMP, possibly through an inhibition of phosphodiesterase activity (Vinggaard et al., 2000). The aromatase enzyme is becoming increasingly important because it is a key target site for the treatment of E2-sensitive mammary tumors in humans.

12.5.4.8  Imazalil In a two-generation reproduction study (1 litter per generation) in Wistar rats given 0, 5, 20, and 80 mg/kg/day for two generations until weaning of the second generation, increased pup/litter mortality, decreased number of litters, and mean litter size were observed. A significant decrease in mean body weights and body weight gain for P0 males during precohabitation and P0 females during pregnancy, birth, and lactation at 80 mg/kg/day resulted in parental NOAEL of 5 mg/kg (females and males in P0 only). The P0 and F1 generation showed an increase in duration of gestation at 80 mg/kg/day and a decrease in litter size (P0: 80 mg/kg/day; F1  20 mg/kg/day). A reduction in the number of litters in P0, number of live pups/female and survival rate (day 4–21), and an increase in the number of dead pups/female at 80 mg/kg/day resulted in a reproductive NOAEL of 5 mg/kg/day. In a teratology study at concentrations of 0 (water), 40, 80, or 120 mg/kg given by gavage to 24 mated female Sprague–Dawley rats/group on days 6 through 16 of gestation (day sperm positive  day 1), adverse developmental effects were indicated. These

420

included increased mean resorptions per litter and reduced mean fetal weights in the absence of significant maternal effects yielding a developmental NOAEL of 40 mg/kg. In mated Cobs CD1 mice given imazalil sulphate at 0, 10, 40, 80, and 120 mg/kg during days 6 through 16 of pregnancy by gavage, increased fetal effects, and embryotoxicity in the absence of equally severe maternal toxicity at 40 mg/kg were noted, yielding a developmental NOAEL of 10 mg/kg/day (CDPR, 1995).

12.5.4.9  Propioconazole In a teratology (Segment II) study in Crl:COBS CD(SD)BR rats, dosed 0, 30, 90, 360/300 mg/kg/day by oral gavage from gestational days 6 to 15 with propioconazole, developmental toxicity included cleft palate, short or absent renal papilla, and dilated ureters. The developmental NOAEL was 30 mg/kg/day. Maternal effects included ataxia, lethargy, and salivation with a maternal NOAEL of 90 mg/kg/day (CDPR, 1990).

12.5.4.10  Maneb Maneb produced fetal hydrocephalus in litters of rats receiving 480 mg/kg/day (Chernoff et al., 1979a). In FIFRA studies, adverse developmental effects appear to have occurred because of contamination of maneb with ethylene thiourea. (CDPR, 2009). The teratogenicity of a commercial formulation of the fungicide maneb (Maneb 80, containing 80% manganese ethylenebisdithiocarbamate and 20% inert ingredients) was evaluated in chick embryos. It was found to be teratogenic at all concentrations tested (0.5, 1.5, 4.5, or 13.5 g/liter maneb aqueous solutions for 30 s), producing mainly unilateral lower limb deformities. No adverse effects on development were noticed after exposure to the inert ingredients (Maci and Arias, 1987).

12.5.4.11  Metam Sodium Adverse effects on the reproductive system were not observed in a rat study, however histopathology in the nasal cavity demonstrated: Bowman’s duct hypertrophy with loss of alveolar cells, disorganization/degeneration/atrophy of olfactory epithelium, hyperplasia of olfactory epithelium and dilatation of ducts of Bowman’s glands, at the high dose in females of both F0 and F1 generations (0.1 mg/ml in the drinking water). Developmental effects were noted in several studies and included severe malformations (meningocele, anophthalmia, hydrocephaly) at 60 mg/kg in the rat. The rat developmental NOAEL  5 mg/kg based on a decrease in fetal weights, numerous skeletal developmental delays at 20 mg/kg and higher levels, and delayed ossification in hand and foot bones at 60 mg/kg. In the rabbit, postimplantation loss, early intrauterine deaths, and total litter resorptions were increased at 60 mg/kg with a developmental

Hayes’ Handbook of Pesticide Toxicology

NOAEL  5 mg/kg/day (based on decrease in mean live litter size, mean litter and fetal weights, and proportion of males/females at 60 mg/kg). Increase in severe defects (cleft palate and meningocele) at 60 mg/kg, and skeletal variations at  20 mg/kg /day were also noted. In another study in Himalayan rabbits embryotoxicity in the form of a statistically significant dose-related increase in dead implantations per pregnant animal at the mid and high dose was noted. At the high dose two fetuses with a neural tube closure defect (meningocele  spina bifida) were observed. This study was also acceptable under FIFRA guidelines; the developmental NOAEL was 10 mg/kg/day based on the increase in dead implantations at 30 mg/kg/day and the maternal NOAEL was 30 mg/kg/day based on decreased food consumption at 100 mg/kg/day (CDPR, 2009).

12.5.4.12  Methyl Thiophanate The fungicide methyl thiophanate widely used to control some of the most common fungal diseases in crops is metabolized in animals into benzimidazole compounds, including the reproductive toxicant carbendazim. However, standard toxicological tests did not indicate that methyl thiophanate may cause testicular toxicity and/or embryotoxicity, which are typical effects of many benzimidazoles. In the B6C3F1 mouse in spite of the high doses administered, none of the testicular parameters examined (sperm head count, specific enzyme activities, histopathology on days 3–35 post-dosing) showed significant alterations as compared to controls at any time post-dosing. Pregnant CD rat dams administered orally the limit dose of 650 mg/ kg BW/day during preimplantation (gestational day or GD 2–5) or peri-implantation (GD 6–9) phases showed maternal toxicity, with only marginal reductions of the growth of embryos and adnexa (Traina et al., 1998). Earlier studies submitted by Atochem North America, Inc. (1985) for methyl thiophanate, show for the rat a teratogenic NOAEL  2500 ppm (125 mg/kg/day) based on the high dose tested (HDT) with a maternal NOAEL  250 ppm (12.5 mg/kg/day), LEL  1200 ppm (60 mg/kg/day) and a fetotoxic NOAEL  2500 ppm (HDT). For mice (study by Pennwalt Corp.) the 1000 mg/kg/day dose caused a decreased number of implantations; other details were unavailable since fetal examinations did not appear to include soft tissue examinations. A three-generation reproduction study in the rat (study by Penwalt Corp.) demonstrated a Reproductive NOAEL  160 ppm (8 mg/ kg/day); LEL  640 ppm (32 mg/kg/day; HDT) based on reduced litter weights (U.S. EPA., IRIS database http:// www.epa.gov/ngispgm3/iris/subst/0336.htm).

12.5.4.13  Pentachlorophenol Pentachlorophenol (PCP) is used primarily as a wood preservative. It has been shown to be fetotoxic and teratogenic

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

during early gestation. Commerical PCP is contaminated with chlorinated dioxins and dibenzofurans, tetrachlorophenols and hydroxychlorodiphenyl ethers (Williams, 1982), and these compounds can exert their own effects. Additionally, PCP was reported to be a contaminant in commercial creosote preparations used in wood preservation and may have contributed to its early fetotoxicity. Pentachlorophenol was not teratogenic in rats (Schwetz et al., 1974). In studies submitted for registration, PCP was found to have adverse effects in the rat developmental toxicity study due to the fetal resorptions, decreased fetal weights, ossification delays, and malformations; these findings cannot be assured to result strictly from maternal toxicity (CDPR, 2009). Maternal NOAEL  30 mg/kg/day (body weight and food consumption decrements) and the developmental NOAEL  30 mg/kg/day (increased fetal resorptions, decreased fetal weights, a modest incidence of malformations such as gastroschisis, hydrocephaly, and diaphragmatic hernia judged to be treatment-related, although not statistically significant). There was a significant increase in the incidence of dilated pelves, and of delayed ossification in several areas, and increased mean numbers of thoracic vertebrae and associated increased incidence of 14th ribs. Another study concluded a possible adverse effect because a relatively low developmental toxicity NOAEL was observed in the absence of maternal toxicity. A maternal NOAEL  200 ppm (13 mg/kg/day), based on reduced weight gain, clinical signs such as ringed eye and possibly vaginal hemorrhaging and a developmental effects NOAEL  60 ppm (4 mg/kg/day), based on reduced fetal weights, misshapen centra, and a possibly treatment-related increase in resorptions (significant increase in females with 2 resorptions) was noted (FDA, 1987), indicating a possible adverse effect.

12.5.4.14  Terrazole Decreased live litter size, fetal weight and pup survival (24 h), increased resorptions and malformations were noted at 45 mg/kg/day in rabbits. Developmental NOAEL  15 mg/ kg/day; adverse effects were indicated, even though the NOAELs for maternal toxicity and developmental toxicity are equivalent, the developmental effects were quite marked (total resorptions  31 in the high dose compared to seven in the control, 24-hour survival of 80% in the high dose compared to 99% in control). The increased incidence of malformations in the high dose included tail defects, underdeveloped hind limbs and crossed hind legs was noted in the Toxicology Summary (CDPR, 2009).

12.5.4.15  Vinclozolin This fungicide has the unique claim to be a compound that results in abnormal rodent sex differentiation following exposure during critical stages of life. Effects such as

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hypospadias, ectopic testes, vaginal pouches, agenesis of the ventral prostate, and nipple retention in male rats were commonly observed (Gray et al., 1994). In the FIFRA reproduction study failure of F1 males to acquire normal anatomical and functional male characteristics, marked retardation in neonatal growth and survival at dose levels not commensurately toxic to adults, and lenticular degeneration were the principal possible adverse effects. In the FIFRA developmental toxicity study in rats decreased anogenital distance in males, a finding which was repeated in all the studies conducted and interpreted as feminization of male fetuses, was observed. Similar findings were not noted in the mouse or rabbit (CDPR, 2009). Vinclozolin administered by dermal application (100 mg/kg in 100 l of dimethylsulfoxide) to rabbits in the peripubertal phase resulted in a smaller weight gain during pubertal growth, and at maturity, the accessory sex glands of the exposed animals weighed less than those of the controls. However, the pooled sperm count of the exposed animals was significantly higher (p  0.017) than that of the unexposed animals, probably because the anti-androgenic effects of vinclozolin may have blocked the negative feedback mechanism of testosterone on the hypothalamus or pituitary gland, allowing for an increase in gonadotrophin release, and consequently increasing sperm production at puberty (Moorman et al., 2000). Exposure to vinclozolin in the diet was shown to cause subtle alterations in locomotor activity and consumption of saccharin-flavored solution with effects more pronounced in females (Flynn et al., 2001).

12.5.5  Rodenticides Reviews on the effects of warfarin exposure indicate an uncommon but strikingly similar pattern of congenital anomalies in children born to women exposed to the compound. The syndrome consists of nasal hypoplasia, stipled epiphyses and growth, retinal-optic atrophy, and central nervous system anomalies (Friedman and Polifka, 1994). While warfarin is used as a rodenticide, it may also be administered to women with heart valve prosthesis. Ginsburg and Baron (1994) recommended not giving warfarin to women between 6 and 12 weeks of gestation.

12.5.6  Animal Health Products,   Fumigants, and Miscellaneous   Pesticides (Tables 12.6 and 12.7) 12.5.6.1  Anthelmintics Several case reports have been published associating anthelmintic drugs with the induction of birth defects in the human. Ectromelia in infants whose mothers were treated with a tin-based taenifuge, multiple malformations (brain, jaw, ear, limb and heart defects) subsequent to

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Table 12.6  Developmental and Reproductive Toxicity Profile of Miscellaneous Pesticides Chemicala

Species

Toxicity profile

Dose (mg/kg)b

Acrylonitrile

Rat

Teratogenic Testicular toxin

25

Murray et al. (1978a,b) [1]

Hamster

Teratogenic

80

Willhite et al. (1981)

Arsenic trioxide

Rat

– Reproduction

10

Benzenesulfonic acid Hydrazide

Mouse

Embryo/fetal mortality

5.5

Benzylbenzoate

Rat

–

1%

Busan 77

Rabbit

–

125

Drake et al. (1990)

Chlorofebrifugine

Rat

–

9.3 (p.o.) or 6 ppm (diet)

Kaemmerer and Seidler (1976) [1]

Chloropicrin

Rat

Reduced fetal weight

3.5 ppm

Inhalation route

Rabbit

Reduced fetal weight

2 ppm

Inhalation route

Chlorosil

Rat

–

100

Boikova et al. (1981)

Cyclonite

Rat

– Reproduction

50

Minor et al. (1982)

Rabbit

–

20

Dikurin

Rat

–

20

Shepelskaya (1988)

Diphenyl

Rat

–

500

Khera et al. (1979a)

Diphenylamine

Rat

Renal lesions

1.50%

Reproduction Ethylene oxide

Mouse

Rat

Comments

Teratogenic by i.p. route 4 generations

References

Stump et al. (1998b) Morris et al. (1938) [1] Matschke and Fagerstone (1977a) [1]

Diet

Morita et al. (1981)

York et al. (1994)

Diet

Crocker et al. (1972)

2-G study

Thomas et al. (1967) [1]

i.v. route, also findings by inhalation route

Kimmel and LaBorde (1979) [1]

Developmentally toxic

150

Teratogenic

150

Developmentally toxic

10 ppm

Inhalation route

Snellings et al. (1982a,b) [1]

i.v. route

Kimmel et al. (1982)

Rutledge and Generoso (1989)

Reproduction Rabbit

Embryotoxicity

9

Gliftor

Mouse

Reduced fertility

300

Tattar (1973) [1]

Guanylthiourea

Rat

Teratogenic

33

Schardein (2000)

Methoxychlor

Rat

Fetopathic

100

Khera et al. (1978)

Cow

–

9.9

Macklin and Ribelin (1971) [1]

Mouse

–

250

Rat

– Testicular toxicity in reproductive studies

70

Methyl bromide

Inhalation route 2-G studies

Hardin et al. (1981) Kato et al. (1986)

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

423

Table 12.6  (Continued) Chemicala

Species

Toxicity profile

Dose (mg/kg)b

Comments

References Hurtt and Working (1988) Eustis et al. (1988)

Rabbit

Agenesis of gall bladder

70

Rat

– Reproduction

15

Rabbit

–

1.5

N-methyl-N-(1naphthyl) fluoro acetamide

Mouse

Fetal growth retardation Teratogenic

20 20

Makita et al. (1970a) [1]

PCA

Rat

Developmentally toxic

4

Welsh et al. (1985)

Peropal

Rat

–

30

King (1981)

Potassium cyanide

Rat

–

500 ppm

Potassium dimethylthiocarbamate

Rat

Fetotoxic

150

Rabbit

Fetotoxic

76

Sulfurylfluoride

Rat

Methylisothiozolinone

Tetramethyl thiodiphenylene phosphorothioate

225 ppm

Rabbit

Decreased fetal weight

225 ppm

Rabbit

–

164

Inhalation route

Hardin et al. (1981) U.S. EPA (1998d)

Diet

Terve and Maner (1981) Drake et al. (1989)

Inhalation route

Hanley et al. (1989)

Or dermal

[2]

Notes: Many of the citations provided in the References column have been taken from the textbook Chemically Induced Birth Defects by Schardein (1993, 2000). These references are indicated by either [1] or [2] following the citation, for Schardein (1993), and Schardein (2000), respectively. All CDPR citations may be accessed at the departmental webpage at http://www.cdpr.ca.gov/docs/risk/toxsums/toxsumlist.htm. ‘Developmental toxicity’ includes reduced fetal weight, increased embryo/fetal mortality and/or increased developmental variations, but does not include malformations. Teratogenic is the term used to indicate presence of malformations. a Includes fumigants, miticides, rodenticides, pediculicides, coccidiostats, molluscicides. b Doses are oral unless stated otherwise; is the LOAEL wherever effects were observed or NOAEL when there were no effects.

Table 12.7  Animal Health Pesticides (Veterinary Antiparasiticalsa) Chemical

Species

Toxicity profile

Dose (mg/kg)b

Amitraz

Rat

Developmental and behavioral changes

20

Palermo-Neto et al. (1994)

Bromofenofos

Rat

Developmentally toxic Teratogenic

50 50

Yoshimura (1987c) [1]

Cambendazole

Rat

Embryotoxic Teratogenic

7.6 7.6

Delatour and Richard (1976)

Sheep

Embryotoxic Teratogenic

50 50

Horse

Teratogenic

20

Schardein (1993)

Carbon tetrachloride

Rat

–

75

Narotsky et al. (1997)

Carbothion

Rat

Fetal death

90

Schardein (1993)

Comments

References

(Continued )

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424

Table 12.7  (Continued) Chemical

Species

Toxicity profile

Dose (mg/kg)b

Crufomate

Rat

– Testicular toxin

500 ppm 1000 ppm

McCollister et al. (1965) [1]

Dibromochloropropane

Rat

Developmentally toxic

25

Ruddick and Newsome (1979) [1]

Rabbit

Reproductively toxic

0.1 ppm

Mouse

–

60

Schwetz et al. (1979b) [1]

Rat

–

25

Schardein (1993)

Rabbit

–

62

Vogin et al. (1971) [1]

Pig

– Reproduction

8.5

Wrathall et al. (1980) [1] Batte et al. (1969)

Cow

–

6.2

Macklin and Ribelin (1971) [1]

Rat

–

100

Fraser (1972) [1]

Rabbit

–

200

Dog

–

2  use level

Over 2 generations

Rodwell et al. (1986a) [1]

Diethylcarbamazine and oxibendazole

Dog

–

13.2/10

Over 2 generations

Rodwell et al. (1987)

Fenbendazole

Rat

–

120

Delatour and Lapras [2]

Flubendazole

Rat

Developmentally toxic Teratogenic

40 40

Yoshimura (1987) [1]

Ivermectin

Rat

Neonatal mortality and decreased pup growth

0.4

Multi-generation study

Lankas et al. (1989) [1]

Primate

–

100 g/kg

In infants

Lankas et al. (1989) [1]

1-generation study

Dichlorvos

Diethylcarbamazine

Comments

Reproductively toxic in man

References

Rao et al. (1982) Whorton and Foliart (1988)

Ivermectin and pyrantel

Dog

–

18/15

Mebendazole

Rat

Embryotoxic Teratogenic

10 10

Delatour and Richard (1976)

Rabbit

–

40

Cited in Shepard (1995)

Naftalofos

Rat

–

15

Kagan et al. (1978) [1]

Netobimin

Rat

Teratogenic

71

Ruberte et al., 1995 cited in Shepard (1995)

Nitroxynil

Sheep

–

34

Lucas (1970) [1]

Oxfendazole

Rat

Teratogenic

16

Delatour et al. (1977)

Sheep

Teratogenic

23

Cow

–

14

Piercy et al. (1979) [1]

Pig

–

13.5

Morgan (1982)

Mouse

–

30

Theodorides et al. (1977) [1]

Rat

–

149

Delatour and Richard (1976)

Oxibendazole

Clark et al. (1992)

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

425

Table 12.7  (Continued) Chemical

Species

Toxicity profile

Dose (mg/kg)b

Parbendazole

Rat

Embryotoxic Teratogenic

10 10

Duncan and Lemon (1974)

Rabbit

Abortion

10

Duncan et al. (1974)

Hamster

–

100

Duncan and Lemon (1974)

Cow

–

60

Miller et al. (1974) [1]

Pig

–

90

Hancock and Poulter (1974) [1]

Sheep

Teratogenic

60

Szabo et al. (1974) [1]

Mouse

–

400

Manufacturer’s information

Rat

– Reproduction

400

Rabbit

–

400

Piperazine

Pig

–

15,000

Ziborov (1982) [1]

PW 16

Rat

Skeletal variations Reproductive effects

440 440

Samojlik et al. (1969b) [1]

Pyrantel

Rat

–

3000

Owaki et al. (1970b) cited in Shepard (1989)

Rabbit

Miscarriage

1000

Owaki et al. (1970a) cited in Shepard (1989)

Horse

–

12.5

Conway et al. (1970)

Sodium arsenate

Mouse

–

120

Terrazole

Rabbit

Increased resorptions and malformations

15

CDPR, Toxicology Summary (1990)

Thiabendazole

Mouse

Teratogenic

700

Ogata et al. (1981) [1]

Rat

Teratogenic

500

Khera et al. (1979b)

200

Szabo et al. (1974) [1]

Permethrin

Sheep

Comments

References

3-G study

i.p./i.v. doses teratogenic in other species

Hood et al. (1972)

Tribendimin

Rat

–

200

Shao et al. (1988) [1]

Triclabendazole

Rat

Reduced fetal growth

100

Yoshimura (1987a) [1]

Notes: Many of the citations provided in the References column have been taken from the textbook Chemically Induced Birth Defects by Schardein (1993, 2000). These references are indicated by either [1] or [2] following the citation, for Schardein (1993), and Schardein (2000), respectively. All CDPR citations may be accessed at the departmental webpage at http://www.cdpr.ca.gov/docs/risk/toxsums/toxsumlist.htm. ‘Developmental toxicity’ includes reduced fetal weight, increased embryo/fetal mortality and/or increased developmental variations, but does not include malformations. Teratogenic is the term used to indicate presence of malformations. a Includes acaracides and anthelmintics. b Doses are oral unless stated otherwise; is the LOAEL wherever effects were observed or NOAEL when there were no effects.

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intake of mebendazole during the first month of pregnancy and the incidence of spina bifida, renal anomalies along with hydrocephalus due to quinacrine administration in the first trimester are noted in a review of the data (Schardein, 1993). Negative reports are also noted for commonly used compounds such as piperazine and ivermectin exposures.

12.5.6.2  Benzimidazole Family of Compounds Variable but teratogenic potential has been noted for the benzamidazoles as a group. As reviewed by Schardein (1993), skeletal abnormalities were reported in sheep and in rats exposed to parbendazole, however teratogenicity was not reported at comparable or higher doses in hamsters, rabbits, cattle, and swine. Cambendazole was noted to have induced multiple defects in the rat and in sheep; flubendazole was found to be developmentally toxic in rats producing multiple malformations; mebendazole induced malformations in rats with up to 100% incidence in a dose-dependent manner but was not teratogenic in rabbits even at high doses. Oxyfenbendazole induced multiple abnormalities in rats and sheep (Schardein, 1993) and swine (Morgan, 1982). Parbendazole has a safety index of over 30 times the recommended dose in healthy animals, but may be teratogenic at doses only slightly higher than the recommended one. It was parbendazole that first alerted scientists to the embryotoxicity of benzimidazoles (Manger, 1991). Thiabendazole is used as a veterinary anthelmintic and fungicide and has variable teratogenic effects in animals. Teratogenicity has been noted in some laboratory animal species (mice and rats), but other studies and reviews have claimed it to be relatively safe (Manger, 1991; Schardein, 1993). In acceptable FIFRA studies submitted for registration, adverse effects were not noted (CDPR, 2009). Another benzamidazole, benomyl (as reported earlier in this chapter), has also demonstrated fetotoxicity and teratogenicity findings in the absence of obvious maternal toxicity indicating possible adverse effects.

12.5.6.3  Antimalarials Chloroquine and its congeners which are inhibitors of dihydrofolate reductase and primaquine are known to exert teratogenic effects, and since they are under the category of prescription drugs, the likelihood of exposure during pregnancy is low. Defects noted after quinine exposure include deafness due to auditory nerve hypoplasia, optidisc problems, limb anomalies and visceral malformations as well as fetal deaths (Schardein, 1993). Rates of spontaneous abortion and birth defects were comparable in pregnant women taking mefloquine (Lariam), compared with chloroquineproguanil, or pyrimethamine-sulfadoxine prophylaxis, in the first trimester of pregnancy (Phillips-Howard and Wood, 1996). Teratogenic effects for mefloquin were observed in animals but data from humans are lacking (Vanhauwere et al., 1998).

Hayes’ Handbook of Pesticide Toxicology

12.5.6.4  Imidacloprid Imidacloprid is widely used against fleas in dogs and cats and also as an insecticide for use on soil, seed or foliar treatment in rice, cereal, vegetables, cotton and turf to control ricehoppers, thrips, termites, turf and soil insects, and some beetle species. In a rat developmental toxicity study, a high percentage of male fetuses and increased incidence of wavy ribs were noted at 94.1 mg/kg/day indicating a possible adverse effect. Maternal NOAEL  25.9 mg/kg/ day (based on decreased body weight gain and reduced food consumption of the 94.1 mg/kg/day treatment group); developmental NOAEL  25.9 mg/kg/day (based on increased incidence of wavy ribs in the fetuses of the 94.1 mg/kg/day treatment group) (CDPR, 2009).

12.5.6.5  Methyl Bromide This gas has been used extensively as a fumigant to combat nematodes in strawberry farming and tomatoes. Alternatives to its use are needed due to its ozone layer depleting properties and it is slated for replacement as per the Montreal Protocol. Chemically it is an alkylating agent and capable of neurotoxicity. The compound appears to demonstrate extreme differences between species, dogs being unable to tolerate doses several-fold lower than those in the rat. Genetic polymorphism for the metabolism of this compound has also been noted. Exposure to methyl bromide in a two-generation reproduction study in Sprague– Dawley rats by inhalation affected fertility (fertility index decreased from 90.9% in the controls to 68% in the 30 and 90 ppm groups) and decreased the body weights of parental and reduced the growth of neonatal rats. Pregnant animals were only exposed 5 days/week (for a total of 14–15 days) during their pregnancy and the pups were not directly exposed until after weaning on postnatal day 28. Parental NOAEL  3 ppm (reduced fertility). Progeny NOAEL  3 ppm based on decreased pup bodyweight and reduced organ weights including reduced F1 brain weight/ reduced width of the cerebral cortex. Data submitted for registration purposes were found to be marginally acceptable, but do not conclusively demonstrate the absence of neurotoxic potential. The developmental study in New Zealand White rabbits demonstrated maternal toxicity at 80 ppm (311 mg/m3) such as reduced body weight and weight gain and clinical signs of central nervous system toxicity. Fetal effects that were not statistically significant but are quite rare, i.e. considered biologically significant, were noted. These include omphalocele, hemorrhaging with or without hydrops, and retroesophageal right subclavian artery. Also gall bladder agenesis, fused sternebrae and decreased fetal body weight were statistically significant at 80 ppm resulting in the NOAEL for maternal toxicity and developmental effects of 40 ppm (155 mg/m3). In rats NOAEL  20 ppm for a developmental toxicity study based on delayed skeletal

Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

ossification and a maternal NOAEL  70 ppm was noted, but the study did not test at a high enough dose level. Reports of other studies via oral exposure in rats and another strain of rabbit demonstrate microphtalmia in rats and some skeletal malformations in the rabbit though not in a dose-responsive pattern (Kaneda et al., 1998). The reports did not meet FIFRA specifications and historical negativecontrol data for the rabbit strain employed (Kbl:JW) are not generally available in the open literature. The oral route provides accurate dosage, but since it is more likely to be metabolized prior to reaching the brain than the inhalation route, it may be argued that a higher dosage may be needed to compare the oral route with the inhalation route. While the pharmacokinetics of transplacental transfer of methyl bromide gas is not available at this time, since methyl bromide is known to have neurotoxic potential, and human exposure is most likely via inhalation or skin, inhalation may be the preferred route to detect neurotoxic damage. Hence adverse effects to development were not observed in the oral studies, but the inhalation studies do demonstrate adverse effects in both reproduction and developmental toxicity studies.

Conclusion Over 4500 chemical tests have been reported using the FDA Segment II protocol and more than one-third have come up as positive for developmental toxicity (J.L. Schardein, personal communication, 1999). Approximately 25–30 chemicals or families of chemicals are considered human teratogens, either on the basis of a Segment II or FIFRA study or by other animal models that have been optimized to detect a toxic effect (Schwetz, 1994). The pilot rangefinding studies in addition to helping to narrow the dose level have also served to be predictive of the definitive study. While much is known about the mode of action of some chemicals, the complexity of development suggests that there may be multiple mechanisms of interference with normal development. These mechanisms are not known for even known human teratogens. Similarly hundreds of chemicals have been tested using protocols for reproductive toxicity. A tenfold less than that for developmental toxicity and one-quarter of these have been documented to be reproductive toxicants to humans and in laboratory animals. There is greater concordance between laboratory animal models and humans for adverse effects on fertility than in the area of developmental effects, e.g. male reproductive toxicants acting on the testes in laboratory animals have the same site of action in humans (Schwetz, 1994). Developmental toxicity in animals, however, does not translate to the same kind of developmental defect in humans (Kimmel et al., 1993; Schardein and Keller, 1989). This lack of concordance as noted for both drugs and other chemicals of commerce has led to the interpretation that some adverse developmental

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effect in an animal study is potentially predictive of some adverse developmental effect in humans. Further confounding the issue are accounts that around 60–70% of pesticides registered for use have not been adequately tested at the laboratory or clinical level (Mott and Snyder, 1987) and that 25% of the compounds exported from the United States are in fact banned or unregistered in the United States (Schardein, 1993). In evaluating global exposure patterns, the data submitted to regulatory agencies become more valuable. Regulatory agencies in the European Community are moving to reduce the number of experimental animals that are being sacrificed for studies. While such a trend is helping to reduce unnecessary wastage of experimental animals, in vivo data submitted to agencies in the United States often serve as critical studies for a specific compound. Regulations in Japan (Ministry for Agriculture, Forests and Fisheries – MAFF) have some similarities to those in the United States and so duplication of studies can occur, but the benefits of such studies serve to reduce the likelihood of a thalidomide disaster. The aim of testing and regulation is thus to minimize the liability to the manufacturers of chemicals, while assuring that the public will be exposed to a safe set of pesticides. This chapter discussed the majority of pesticides that have been used and reviewed the available data on their reproductive and developmental toxicity. Details for a specific compound or new active ingredient may be obtained from the databases that are accessible to the public at state and federal agencies or from publications.

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Chapter | 12  Developmental and Reproductive Toxicology of Pesticides

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Chapter 13

Microbial Pest Control Agents: Use Patterns, Registration Requirements, and Mammalian Toxicity Andrew L. Rubin California Environmental Protection Agency, Sacramento, California

13.1  Introduction Concern for the safety of agricultural workers and the public, as well as for the integrity of ecosystems, has fueled an interest in the use of microbes as pest control agents (Siegel and Shadduck, 1992). In addition to such pest management dividends as low toxicity and low environmental impacts, microbial pest control agents (MPCAs) offer high target selectivity and extended pest control in cases where establishment of the microbe occurs in local habitats (Khetan, 2001; Weinzierl et al., 2005). However, MPCAs have not significantly displaced conventional pesticides in major crop systems. Global MPCA sales in 2001 amounted to $160 million, less than 1% of the $28 billion pesticide market for that year (Zahodiakin, 2002). This may be due to disadvantages that are relatively specific to the use of MPCAs, including lability in the face of environmental stressors (temperature, moisture, ultraviolet radiation, etc.), unique formulation and storage requirements, high expense due to narrow target species ranges and occasional overselectivity, which leaves other pests active in the treated environment (Weinzierl et al., 2005). Grower skepticism resulting from past microbial pest management failures may also play a role in their low acceptance rate (Zahodiakin, 2002). From a regulatory standpoint, MPCAs present special challenges due to their unique properties as living organisms, including their theoretical potential to persist within, and colonize, humans and other mammals. Following the registration in the United States of Bacillus popillae as an insecticide in 1948, the list of federally registered MPCAs has grown to include not only bacteria but also viruses and eukaryotes. There were 245–250 products listing approximately 60 different Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

MPCAs as active ingredients under federal registration in August 2000, including as separate counts different subspecies or strains of the same microbial species (personal communication, R. Torla, U.S. EPA). By March 2007 the number of registered MPCAs had risen above 70, with over 220 products listed [U.S. Environmental Protection Agency (EPA), 2007]. In California, 91 products containing 33 MPCAs were under active registration in September 2000, rising to 103 products containing 47 MPCAs by June 2008 [Department of Pesticide Regulation (DPR) registration database]. The number of pounds of MPCA active ingredients applied in agriculture increased over sixfold in California between 1990 and 2007 (DPR, 2008a), with Bacillus thuringiensis (Bt) products ­ overwhelmingly dominating this segment of the market (Table 13.1). Nonetheless, the use of MPCAs occupying a smaller market fraction, notably Bacillus pumilus, Bacillus sphaericus, Bacillus subtilis, Beauveria bassiana, and Myrothecium verrucaria, also rose during this period. Increasing human exposure to these organisms under both occupational and nonoccupational scenarios is thus a reasonable expectation. In this chapter, the regulatory system in place in the United States for assessing the toxicity, infectivity, and pathogenicity of MPCAs to humans is reviewed. In addition, toxicologic overviews for several prominent or proposed MPCAs are provided. These overviews are directed primarily at toxicity and pathogenicity issues arising from exposure to viable microbial organisms, though individual microbial toxins are considered in some cases.a It is hoped that the reader will gain an appreciation for the unique a Due to the wide-ranging nature of this material, reviews are cited in addition to primary sources in this chapter.

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Table 13.1  Total Pounds of MPCA Active Ingredients Applied per Year in California, 1990–2007a

Agrobacterium radiobacterb

1990

1991

1992

1993

1994

1995

1996

1997

1998

1999

2000

2001

2002

2003

2004

2005

2006

2007

0

0

0

2

4

6

17

37

60

10

2

1

5

3

2

0

10

4

0

3

9

40

4

4

2

0

0

0

0

0

0

Ampelomyces quisqualis

0

Aspergillus flavus, strain AF36 Bacillus pumilus, strain QST 2808

0

2

3,546

5,636

6,970

Bacillus sphaericus, serotype H-5A5B, strain 2362

1,298

4,886

2,274

2,749

7,941

4,667

10,158

14,187

34,154

45,430

20,443

Bacillus subtilis, GB03

0

0

0

0

1

4

5

7

15

14

6

882

7,201

18,969

17,271

16,618

14,038

17,135

16,963

Bacillus subtilis (dried), strain QST 713 Bacillus thuringiensisc

4,552

3,528

Bacillus thuringiensis, subsp. aizawai, serotypes H-7

3,724

1,071

476

1,562

520

182

751

24

76

115

16

11

12

16

35

26

2

802

4,935

8,050

10,182

14,210

10,854

10,422

9,064

5,511

3,886

7,541

3,014

2,335

1,752

2,877

10,540

21,941

27,075

33,336

18,378

32,513

Bacillus thuringiensis, subsp. aizawai, strain ABTS-1857 711

Bacillus thuringiensis, subsp. aizawai, strain GC-91

1,936

5,115

6,529

7,406

4,282

Bacillus thuringiensis, subsp. aizawai, strain SD-1372 2,416 Bacillus thuringiensis, subsp. israelensis, serotype H-14

3,391

6,070

9,236

4,619

6,827

4,613

4,459

13,180

3,017

4,419

3,953

3,980

5,024

4,088

11,255

9,377

20,395

3

158

498

1,322

558

347

315

432

562

5,038

88,039

24,711

8,266

11,376

9,311

11,297

14,394

8,746

271

9,485

29,326

23,001

41,734

59,018

36,326

31,046

3,423

6,161

3,916

1,931

2,272

952

3,021

15,491

38,018

46,754

57,985

53,346

69,913

79

164

130

10

1

3

Bacillus thuringiensis, subsp. israelensis, strain AM 65-52 Bacillus thuringiensis, subsp. kurstaki, serotype 3A, 3B

23,234

26,411

30,099

32,834

39,667

40,104

26,051

30,286

21,683

15,231

14,477

Bacillus thuringiensis, subsp. kurstaki, strain ABTS-351 Bacillus thuringiensis, subsp. kurstaki, strain BMP 123

6

1

33

2,714

Bacillus thuringiensis, subsp. kurstaki, strain EG 2348

3,625

Bacillus thuringiensis, subsp. kurstaki, strain HD-1

3,205

1,467

5,207

2,191

2,140

2,743

1,481

222

107

211

281

147

6

835

21,037

23,588

22,300

17,819

10,654

7,173

4,725

3,185

6,139

2,259

0

0

10,548

13,540

22,282

19,676

20,348

53,051

54,234

63,849

7,375

7,132

23,432

27,118

16,576

16,580

16,402

22,702

139

58

19

39

2

5

1

6,482

14,734

439

1,527

930

1,919

1,384

154

Bacillus thuringiensis, subsp. kurstaki, strain M-200 Bacillus thuringiensis, subsp. 21,210 kurstaki, strain SA-11

15,805

10,035

7,865

6,416

8,645

8,691

11,662

9,616

8,730

Bacillus thuringiensis, subsp. kurstaki, strain SA-12 1,564

Bacillus thuringiensis, subsp. kurstaki, strain 2371

3,327

8,291

7,042

7,466

3,468

2,752

1,633

213

Bacillus thuringiensis, subsp. kurstaki, genetically engineered strain EG7826

0

0

Bacillus thuringiensis, subsp. kurstaki, genetically engineered strain EG7841

257

15,619

12,522

12,830

16,682

8,681

681

1,503

344

338

3,872

632

Bacillus thuringiensis, subsp. kurstaki, endotoxins Cry 1AC and 1C (gen. eng.) encapsulated in killed Pseudomonas fluorescens

3,663

29,895

12,634

8,055

7,166

2,211

258

54

5

3

0

1

Bacillus thuringiensis, subsp. kurstaki encapsulated -endotoxin in killed Pseudomonas fluorescens Bacillus thuringiensis, subsp. san diego Bacillus thuringiensis, subsp. san diego encapsulated -endotoxin in killed Pseudomonas fluorescens

21

35

1,823

7,959

14,341

14,535

31,043

44,554

35,129

28,433

17,792

6,419

2,946

445

114

7

6

32

38

53

44

10

1

3

26

8

34

18

8

1

2

1

0

2

2

7

13

34

1

6

1

6

2

1

914

677

1,040

863

624

570

703

0

0

2

Beauveria bassiana, strain GHA

1

573

1,250

920

Candida oleophila, isolate I-182

414

726

216

55

715

(Continued)

Table 13.1  (Continued) 1990

1991

1992

1993

1994

Codling moth granulosis virus

1995

1996

1997

1998

1999

2000

2001

2002

2003

2004

2005

321

Conothyrium minitans, strain con/m/91-08 Gliocladium virens GL-21 (spores)

15

144

156

104

89

60

314

2006

2007

0

0

103

171

198

6

11

6

110

48

30

19

1

152

58

0

Lagenidium giganteum (California strain)

87

151

0

134

859

499

Metarhizium anisopliae, var. anisopliae, strain ESF1

1

1

0

3

37

15

18

15

22

0

0

0

0

0

1,097

8,496

18,824

20,869

45,917

36,104

47,037

39,789

27,977

25,039

28,616

0

0

0

0

0

0

0

0

0

0

0

1

1

1,971

841

896

1,004

614

20

0

0

Myrothecium verrucaria, strain AARC-0255 Nosema locustae spores

0

0

0

0

1

0

5

Paecilomyces fumosoroseus apopka, strain 97 Polyhedral occlusion bodies of nuclear Helicoverpa zea polyhedrosis virus 0

Pseudomonas fluorescens, strain A506

206

3,044

3,639

3,660

2,083

103

1,102

34

Pseudomonas syringiae, strain ESC-11 Pseudomonas syringiae, strain ESC-10 0

Streptomyces griseoviridis, strain K61

21

15

0

0

1

2

5

1,361 0

0 2

5

2

1

1

0

0

Streptomyces lysicus wvec 108 65

Trichoderma harzianum rifai, strain KRL-AG2

39

60

121

125

116

55

41

37

16

1 0

0

24

38

Total

51,431

50,772

55,145

68,817

82,248

96,658

101,945 171,075 168,241 142,700 232,645 205,847 181,192 255,268 233,334 317,628 346,344 336,607

Bacillus thuringiensisrelated total

51,431

50,772

55,144

68,815

82,156

95,937

98,244

163,372 148,577 117,810 206,914 142,551 118,771 177,846 160,681 236,138 251,469 262,089

a Data are from the Pesticide Use Report (DPR, 2008). The figures represent the amounts applied predominantly under agricultural conditions. Where no figure appears, there was no reported usage. Where “0” appears, it is the result of rounding off very small reported usages. b Includes products in which the strain K1026 is identified and products in which the strain is not identified in the PUR. c A strain or subspecies was not provided in the PUR.

Chapter | 13  Microbial Pest Control Agents: Use Patterns, Registration Requirements, and Mammalian Toxicity

problems confronting regulators as they assess the likelihood of human health impacts resulting from the use of MPCAs.

13.2  Toxicity testing requirements for MPCAs By 1981, with the publication by the World Health Organi­ zation of an approach to the safety testing of MPCAs (WHO, 1981), it was clear that differences between conventional chemicals and MPCAs required the development of a separate MPCA toxicity testing scheme. Shadduck (1983) outlined the premises upon which conventional chemical testing was (and is) based and provided reasons why these were not applicable to MPCAs. These premises, as rearticulated by Siegel (1997), were as follows. First, high doses of chemicals generate biological effects, which are expressed either as overt toxicity or as cellular or organ system responses designed to detoxify and excrete the chemical. With MPCAs it is often impossible to generate such effects without first killing the host by suffocation or by circulatory or gastrointestinal blockage. Second, metabolic and excretion pathways are often helpful in predicting conventional chemical toxicity. However, MPCAs are not known to be degraded or genetically altered during passage through the host. Third, persistence and accumulation of chemicals within host organisms necessitates long-term testing for chronic effects. In general, MPCAs neither colonize mammals nor produce chronic effects. [Two caveats: (a) Some MPCAs are capable of persisting within mammals for longer than a few days without multiplying. This necessitates careful examination of their host clearance pattern, which would allow persistence to be distinguished from active infection (Siegel and Shadduck, 1990a). (b) Viral agents targeted at mammalian pests present unique problems due to their mammalian host ranges (see the discussion of rabbit hemorrhagic disease virus).] Fourth, structure–activity relationships, which are often applicable to conventional chemicals, are irrelevant to MPCAs. The WHO testing scheme for determining the toxicity of MPCAs was based on four principles (WHO, 1981): (1) MPCAs pose “inherently different” risks to humans than conventional pesticides; (2) findings of minimal or no toxicity in laboratory testing (“negative results”) are likely; (3) tiered testing, whereby negative results at one level preclude testing at higher levels, is appropriate; and (4) testing protocols should maximize the possibility of generating adverse effects in the host organism. This practical and health-protective approach allowed for a more expedient registration process than that in effect for conventional chemicals. For most potential agents which show commercial promise, negative results under the short-term Tier I requirements obviate higher tier testing. Furthermore, due

445

to their lack of toxicity and pathogenicity, the U.S. EPA has exempted all MPCAs that have been registered to date from the requirement of a tolerance,b though tolerance exemptions for some MPCAs (e.g., B. thuringiensis and nuclear polyhedrosis virus of Heliothis zea) include manufacturing quality control restrictions to mitigate potential batch contamination and other dietary risks (personal communication, A. Reynolds, U.S. EPA). In 1982, the U.S. EPA published its “Pesticide Assessment Guidelines, Subdivision M: Guidelines for Testing Biorational Pesticides” (U.S. EPA, 1982). This document, which was revised in 1989, incorporated WHO’s philosophy and testing schemes into a series of test guidelines that, with some modification, inform the study requirements still in force today in the United States [see the Code of Federal Regulations, section 40, part 158 (40 CFR 158)]. A brief discussion of these guidelines and requirements is presented in the following paragraphs. Other treatments can be found in the literature (Betz et al., 1990; McClintock, 1999; McClintock et al., 1995; Siegel, 1997; Siegel and Shadduck, 1992).

13.2.1  Tier I Under 40 CFR 158, Tier I, acute toxicity/pathogenicity testing by three systemic exposure routes – oral [U.S. EPA Office of Prevention, Pesticides and Toxic Substances (OPPTS) guideline #885.3050], pulmonary (OPPTS #885.3150), and injection (intravenous or intraperitoneal, OPPTS #885.3200c) – is required for the registration of MPCA technical grade active ingredients. Intratracheal or intranasal instillation is often used to fulfill the pulmonary requirement, as aerosolization of viable microorganisms for inhalation studies can be problematic. Both the injection and the intratracheal routes are more invasive than the exposure routes expected in the field, satisfying the WHO principle of maximizing the likelihood of adverse effects. A limit dose of 107 (injection) or 108 (oral and pulmonary) colony forming units (cfu) or the highest obtainable dose, is administered to mice or rats. The animals – three/sex for acute oral and injection, five/sex for acute pulmonary – are monitored over a 4-week period for mortality, clinical signs, body weight, gross pathology, and microbial clearance, a measure of the ability of the host to remove invading microorganisms over time (thus an indication of the presence or absence of an active infection) (McClintock et al., 1995). Clearance is usually assessed by culturing the b

A tolerance is defined as the allowable residue level in food crops for pesticides registered for use in those crops. It is established for each pesticide through analysis of residue and toxicity data submitted under the Federal Insecticide, Fungicide and Rodenticide Act. c  The injection route is not required for viral MPCAs. Intraperitoneal injection is recommended when test material “size or consistency may prevent use of an intravenous injection” (40 CFR 158).

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MPCA from homogenates of various organ systems, body fluids, and excretory products at established time intervals after dosing. Colony forming units are enumerated in vitro and a pattern of clearance is established. While complete clearance can be demonstrated within a few days for most MPCAs (a result not unexpected in view of the fact that these organisms are rarely adapted for life under mammalian body conditions), the process can take 50 days or more for certain persistent organisms. In such cases it is sufficient to demonstrate a clearance pattern and to show that the organism does not produce an active infection which can colonize and multiply within the host. An additional set of Tier I acute toxicity tests by the oral (OPPTS #870.1100), dermal (OPPTS #870.1200), and inhalation routes (OPPTS #870.1300) is required for the manufacturing-use and end-use products. These productoriented tests are similar to those required for conventional chemicals and entertain the possibility that formulation ingredients other than the microbial active ingredient could be hazardous. A limit dose approach is recommended, by which 2 g/kg body weight for oral and dermal toxicity and 2 mg/l for inhalation toxicity is administered to five animals/sex and survival monitored for 2 weeks. For the acute dermal study, the guidelines specify that the test article be applied to the clipped or shaved dorsal or ventral surface of adult albino rabbits. Local irritation as well as systemic toxicity can be evaluated by this approach if a separate primary dermal irritation study (described later) is not carried out. The test article, which if solid is moistened with physiological saline, is applied to the skin under porous gauze for 24 h, after which it is washed off. The 4-h inhalation test is invoked only “when the product consists of, or under conditions of use would result in, an inhalable material” (40 CFR 158). With respect to all of these studies, the U.S. EPA occasionally allows potential product registrants to use toxicity data on the technical grade active ingredient to satisfy the manufacturing-use and end-use toxicity requirements if the formulations are similar. Irritation studies by the dermal and ocular routes (OPPTS #870.2500 and 870.2400, respectively) are also required for manufacturing-use and end-use products. This is presumably because formulation components may be drivers of irritation reactions caused by MPCA products, as is sometimes the case with conventional chemicals. The dermal irritation test involves the application of 0.5 ml of a liquid test substance or 0.5 g of a moistened solid test substance to a shaved area (6 cm2) on the dorsal trunk surface of an adult albino rabbit, three per test (variations apply when the test material is corrosive). Each application site is covered with gauze. The exposure duration is 4 h. After removal of the gauze, edema and erythema reactions are gauged at 0.5–1, 24, 48, and 72 h post application. Evaluations continue for as long as 14 days if dermal irritation is persistent. For the ocular irritation test, 0.1 ml of a liquid test article or 0.1 g of a solid test

article is placed in the conjunctival sac of one eye of each of three rabbits (or one rabbit if severe irritation is anticipated), with the untreated eye serving as a control in each case. Rinsing of the treated eye should not be done for at least 24 h after application. Examinations are conducted at 1, 24, 48, and 72 h post dose, continuing further if irritation persists. Clearance determinations are not required for dermal or ocular exposures. Nonetheless, putative MPCAs occasionally do establish themselves at least temporarily in the eyes. For example, ocular applications of B. sphaericus and B. thuringiensis subsp. israelensis in the rabbit eye led to detections at that site for as long as 8 weeks post dosing (Siegel and Shadduck, 1990a). Cell culture tests are required only in the case of viral pest control agents because other classes of MPCAs are not likely to initiate infections of individual cells. A number of tests using both primary mammalian cell cultures and established mammalian cell lines are necessary to evaluate the toxicity and infectivity of the form of the virus considered to be most infective in susceptible cell cultures or in whole organisms (e.g., insects). These include a plating efficiency test, an infectivity evaluation, and a cell morphologic transformation assay. The latter assay, done specifically in Syrian hamster embryo cells, would be extended to an examination of viral tumorigenicity in hamsters if the MPCA proved capable of morphologically transforming cells in culture. Finally, Tier I requires that reports be filed of hypersensitivity incidents occurring during the manufacture, testing, or use of MPCAs or MPCA products.

13.2.2  Tier II An additional acute oral study (OPPTS #885.3500) may be required under Tier II “when significant toxicity in the absence of pathogenicity and significant infectivity is observed in acute oral, injection, or pulmonary studies (Tier I)” (40 CFR 158). The routes that produced such effects in Tier I testing are used to establish the median lethal dose and slope after a 14-day post dose observation period. If, on the other hand, “significant infectivity and/or unusual persistence is observed in the absence of pathogenicity or toxicity,” a subchronic study by the oral or pulmonary route may be required (OPPTS 885.3600). In a subchronic study the test article is administered daily for at least 90 days at a dose of at least 108 cfu/ animal/day. The animals are monitored throughout for toxicity and pathogenicity/infectivity, with organs, tissues, and body fluids assayed for the presence of the microorganism.

13.2.3  Tier III Tier III testing may be necessary if there are findings of significant toxicity, pathogenicity, or infectivity in Tier I or II

Chapter | 13  Microbial Pest Control Agents: Use Patterns, Registration Requirements, and Mammalian Toxicity

studies. Tier III studies include reproductive fertility effects (OPPTS #885.3659; triggered by evidence of viral persistence or replication in cell culture, inamenability of the MPCA to classification even while it is known to be related to microbes parasitic to mammalian cells, or the presence of a contaminant that is parasitic to mammals), carcinogenicity (OPPTS #870.4200; triggered “for products known to contain or suspected to contain carcinogenic viruses or for microbial components that are identified as having significant toxicity in Tier II testing”), immunotoxicity (OPPTS #870.7800; triggered “for products known to contain or suspected to contain viruses that can interact in an adverse manner with components of the mammalian immune system”), and infectivity/pathogenicity analysis (OPPTS #885.3000; triggered “for products known to contain or suspected of containing intracellular parasites of mammalian cells, for products that exhibit pathogenic characteristics in Tier I and/or Tier II, for products which are closely related to known human pathogens based on the product analysis data, or for known human pathogens that have been ‘disarmed’ or rendered nonpathogenic for humans”). Organisms targeted against mammalian pests clearly present a conundrum to this testing scheme, though it appears that none have yet been considered for registration in the United States.

13.3  Toxicity of individual MPCAs The following discussion focuses on specific toxicity issues pertaining to selected microbes which are either in use or have been considered for use as MPCAs.

13.3.1  Bacteria 13.3.1.1  Bacillus thuringiensis B. thuringiensis is the best known and most widely used of all pesticidal microbes. This gram-positive, spore-­forming, facultative soil saprophyte was first isolated in 1901 by I. Shigetane from infected silkworm larvae and again in 1911 by E. Berliner from diseased flour moth larvae found in a Thuringian mill. The insecticidal activity of the crystalline parasporal protein inclusions was elucidated in 1954 by T. Angus (Angus, 1954). Commercialization of B. thuringiensis occurred first in France in the 1930s, though registration in the United States was not achieved until 1961 (U.S. EPA, 1998a), 3 years after a temporary tolerance exemption was granted for use in food and forage crops (Fisher and Rosner, 1959). Expression of insecticidal genes in tobacco and tomato plants followed in 1987; by 2001, transgenic cotton, corn, and soybeans were dominant forces in the U.S. market, comprising 69, 26, and 68%, respectively, of the total crop (Nester et al., 2002). B. thuringiensis subspecies have shown specificity against various orders of insects including lepidopterans

447

(B. thuringiensis subsp. kurstaki, morrisoni, and aizawai), dipterans (B. thuringiensis subsp. israelensis; Figure 13.1), and coleopterans (B. thuringiensis subsp. tenebrionis). More recent B. thuringiensis isolates are active against nematodes, mites, and protozoa, as well as against other insect orders (Schnepf et al., 1998). The entomopathogenic activity is primarily based on production during the stationary growth phase of a parasporal protein crystal. The crystal is composed of “Cry” (for “crystal”) and – at least in B. thuringiensis subsp. israelensis and morrisoni (among currently commercially relevant B. thuringiensis subspecies) – “Cyt” (for “cytolytic”) proteins. Knowledge of the identity, specificity, and structure of these “-endotoxins” has expanded enormously over the past three decades, with coding sequences known for over 100 of these ­primarily plasmidlocalized genes (Schnepf et al., 1998). The Cry protoxin is activated by solubilization and proteolytic cleaveage under the alkaline gut conditions prevalent in susceptible insects. Cry-based toxicity is precluded under the acidic gut conditions present in most mammals (including humans). The activated protein causes larval death by ­ receptormediated lysis of the midgut epithelium (McClintock et al., 1995; Schnepf et al., 1998). Most Cry genes code for proteins in the 65- to 138-kDa range, with size at least partially dependent on the strain pathotype (Beegle and Yamamoto, 1992; Drobniewski, 1994). Differences in insect toxicity may be a function of different Cry solubilities in the insect gut, as well as different inherent characteristics such as receptor affinity (Schnepf et al., 1998). In some cases the spores can contribute to the insecticidal activity of the parasporal crystal proteins, perhaps through vegetative growth and the resultant production of other toxins (Beegle and Yamamoto, 1992; Nester et al., 2002). The Cyt toxins are hemolytic and cytolytic proteins with protoxin molecular weights in the 25- to 28-kDa range (Drobniewski, 1994). Cyt proteins do not exhibit sequence homology with Cry proteins (Hofte and Whiteley, 1989). They appear to disrupt insect cell membranes through detergent-like effects (Butko et al., 1997) and/or through the formation of cation-selective channels (Drobniewski, 1994). Thomas and Ellar (1983) found that intravenous injection of solubilized parasporal crystal proteins from B. thuringiensis subsp. israelensis was toxic to mice, in contrast to a lack of toxicity upon injection of a similar preparation from B. thuringiensis subsp. kurstaki. This was probably due to the presence of a 28-kDa Cyt protein in the former preparation. The toxicity of the isolated 28-kDa protein from B. thuringiensis subsp. israelensis was subsequently verified by intraperitoneal injection into mice (Mayes et al.,1989). Interestingly, neither the solubilized kurstaki nor israelensis preparations provoked a toxic response in mice by the oral route (Thomas and Ellar, 1983). Several other B. thuringiensis molecules deserve mention. The -exotoxin (thuringiensin), a heat-tolerant adenine

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Figure 13.1  Sporulating cell of Bacillus thuringiensis subsp. israelensis and parasporal bodies characteristic of this subspecies as revealed by transmission electron microscopy. (A) Sporulating cell illustrating the developing spore (Sp) and parasporal body. The parasporal body (PB) is composed primarily of four insecticidal proteins, Cry4A, Cry4B, Cry11A, and Cyt1A. These are assembled during sporulation outside the exosporium membrane (E). (B) Portion of sporulating cell just prior to lysis. The Cry11A crystal (*) lies adjacent to the Cyt1A and Cry4A and Cry4B inclusions. (C) Purified parasporal body showing its components. In this subspecies, the individual protein inclusions are enveloped in a multilamellar fibrous matrix (arrowheads) of unknown composition, which also surrounds the crystals holding them together. A typical mature parasporal body of this subspecies measures 500–700 nm in diameter. Bar in A  100 nm. (Photo and caption courtesy of Brian A. Federici, University of California, Riverside.)

nucleotide toxic to house flies, mammals, and other nontarget organisms, has been demonstrated in B. thuringiensis subsp. thuringiensis and in one B. thuringiensis subsp aizawai strain (Khetan, 2001; U.S. EPA, 1998a). -Exotoxin operates by inhibiting DNA-dependent RNA polymerase (Beegle and Yamamoto, 1992; Sebesta and Horska, 1970). For purposes of registration it is necessary to demonstrate the absence of -exotoxin in B. thuringiensis formulations (McClintock et al., 1995). In addition, a proteinaceous, heat-labile, insecticidal -exotoxin with a molecular weight in the 45- to 50-kDa range has been identified (Beegle and Yamamoto, 1992). -Exotoxin has properties similar to the 50-kDa enterotoxin of B. cereus. Finally, insecticidal activity can be enhanced by the expression of other proteinaceous toxins, among them phospholipases, proteases, chitinases, zwittermicin, and secreted vegetative insecticidal proteins (Khetan, 2001; Schnepf et al., 1998). Extensive testing of intact commercial B. thuringiensis strains has not resulted in appreciable toxicity, pathogenicity, or infectivity (U.S. EPA, 1998a). Siegel (2001) reviewed the toxicity and infectivity of various B. thuringiensis strains, concluding that commercial preparations posed essentially no risk to human populations. One early study cited by Siegel showed that ingestion by humans of

3  109 B. thuringiensis spores/day (the subspecies was not identified) for 5 days, or by rats of 2  1012 spores/kg, produced no toxicity (Fisher and Rosner, 1959). Similar indications of low or no toxicity were obtained in a series of acute studies conducted between 1973 and 1999 by the oral, dietary, inhalation, intranasal, intraperitoneal, intratracheal, intracerebral, and ocular routes in humans, mice, rats, rabbits, and sheep using primarily B. thuringiensis subsp. kurstaki and israelensis (cited in Siegel, 2001). More recent studies have confirmed these results. Using a wettable powder formulation of B. thuringiensis subsp. kenyae, Meher et al. (2002) found that neither oral administration of up to 1.25  108 spores nor dermal administration of 2.5  107 spores to HA rats resulted in systemic toxicity through 3 weeks post dose, though minor erythema was observed within the first 3 days in the dermal study. In addition, conjunctival instillation of 2.5  106 spores did not elicit ocular irritation within 14 days. Toxic reactions are possible when noncommercial subspecies, unusual exposure routes, or, as indicated previously, the use of isolated toxins as opposed to whole organisms are examined. Intracerebral exposure of laboratory rats was lethal when sufficient numbers of organisms were injected (Siegel and Shadduck, 1990b), though human exposure by that route is certainly unlikely. Warren et al. (1984) reported an incident

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in which local and lymphatic inflammation requiring antibiotic therapy occurred when a laboratory worker sustained an accidental injection with spent medium containing B. thuringiensis subsp. israelensis and Acenitobacter calcoaceticus var. anitratus. In that case, A. calcoaceticus, a skin-dwelling bacterium, may have provided the proteases necessary to activate the protoxin by releasing them either into the spent medium or into extracellular fluids at the injection site. It was, nonetheless, unclear to what extent the pathology was due to intoxication and to what extent to bacterial persistence or infection with an accompanying inflammatory response from the host. Isolated health concerns pertaining to noncommercial B. thuringiensis strains have occasionally surfaced. B. thuringiensis subsp. konkukian (serotype H34) was detected in the wounds of a French soldier injured by a land mine explosion (Hernandez et al., 1998). The ability of this strain to cause tissue damage was demonstrated by cutaneous application of bacterial suspensions isolated from the wounds to normal and immunosuppressed mice. Inflammatory lesions developed in all mice treated with 107 cfu. These healed spontaneously in the normal animals but progressed in the immunosuppressed animals. In a study reported in the Russian literature, B. thuringiensis subsp. galleriae was shown to cause syndromes in humans similar to those found in B. cereus-related food poisoning (Pivovarov et al., 1977). Despite the absence of mammalian toxicity in in vivo testing by most economically important B. thuringiensis strains, the ability of these bacteria to persist for extended periods within mammals after injection or intratracheal administration (McClintock et al., 1995; Siegel and Shadduck, 1990b) has occasioned concern, particularly in light of the close phylogenetic relationships between B. thuringiensis and other medically significant Bacillus species (especially B. cereus, B. anthracis, and B. sphaericus). B. thuringiensis with cytotoxic characteristics similar to enterotoxin-producing B. cereus was isolated from stools in a gastroenteritis outbreak in a Canadian chronic care institution (Jackson et al., 1995). While the B. thuringiensis subspecies was not identified, the expression of B. cereus traits was not unexpected, as B. thuringiensis and B. cereus are variants of the same species (Schnepf et al., 1998). Nonetheless, Siegel (2001) felt that the pathogen in this case was more likely to be the Norwalk virus, which was also present in the cohort examined. The production of diarrheal enterotoxin was demonstrated in various commercial preparations of B. thuringiensis by Damgaard (1995), though the levels were generally low compared to those found in a reference culture of B. cereus that had been isolated from a food poisoning outbreak. It was noted, however, that a role for B. thuringiensis in food poisoning may be underestimated due to the need for a special staining technique to differentiate B. thuringiensis from the more conventionally assayed B. cereus. Bishop

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et al. (1999) detected B. cereus-type enterotoxin in six strains of B. thuringiensis (three of long-term commercial importance), though rats experimentally exposed to those strains were unaffected. In addition, -exotoxin was detected in the three noncommercial strains. The authors speculated that the rat system may be not be appropriate for B. cereus-type enterotoxin testing. Using a cell culture assay, Tayabali and Selegi (2000) confirmed the potential of both the kurstaki and the israelensis subspecies to produce nonspecific toxins similar to those found in B. cereus strains. However, the possibility of systemic pathology mediated by those commercial strains was considered unlikely in humans due to the need for sustained infection to produce vegetative cells and cytolytic exoproducts in sufficiently high quantities. Jensen et al. (2002) detected B. thuringiensis subsp. israelensis in stool samples from 8 of 20 subjects exposed during a greenhouse spraying operation with that subspecies. However, there was no correlation in this very small study between the positive detections and any adverse gastrointestinal symptoms. Yang et al. (2003) detected the B. cereus enterotoxin hemolysin BL in 59 B. thuringiensis strains and in five isolates from commercial B. thuringiensis products, and they demonstrated the toxicity of all of these strains in Chinese hamster ovary cells. Nonetheless, Siegel (2001) maintained that there is little direct evidence for B. cereus-type food poisoning outbreaks resulting from B. thuringiensis exposures, listing two possible reasons for this conundrum: (1) Commercial B. thuringiensis fermentation processes may result in products lacking the requisite toxins, an observation supported by the lack of toxicity in numerous studies, including epidemiologic studies in exposed human populations; and (2) B. cereus and B. thuringiensis are not properly distinguished in conventional tests conducted by food safety laboratories and hospital clinics, resulting in an underestimate of the role of B. thuringiensis in incidents of food poisoning. Under field conditions, reports of clinically significant symptoms in humans are rare considering the length of time that B. thuringiensis has been in use as a pesticide. In one case, a farmer developed a corneal ulcer containing B. thuringiensis after being splashed in the eye with a commercial B. thuringiensis product (Samples and Buettner, 1983). A survey of farm workers exposed to commercial B. thuringiensis subsp. kurstaki sprays failed to identify clinical syndromes in the eye, respiratory tract, or skin (Bernstein et al., 1999). However, positive skin-prick allergy tests and induction of IgG and IgE antibodies were documented in some exposed individuals, suggesting that allergenicity could result from repetitive exposure. The possibility of IgE-mediated sensitization also arose in a multiyear study of 329 Danish greenhouse workers exposed to B. thuringiensis products; 23–29% of the sera showed evidence of B. thuringiensis-stimluated IgE in the second year of the study, with 86–88% concordance between the second and third year results (Doekes et al., 2004).

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Two epidemiologic studies examined the consequences of exposure of the general public to B. thuringiensis subsp. kurstaki. While minor reactions were occasionally evident, serious illnesses were not apparent over the course of the studies: 1. Noble et al. (1992) examined records for 26,000 “hotline” phone calls, 1140 family practice patient visits, and 3500 emergency room admissions, in addition to monitoring 120 ground spray workers, during and after a 1992 combined aerial and ground spraying operation with B. thuringiensis subsp. kurstaki to control Asian gypsy moths in southern British Colombia. Of the 19,893 calls received during the 2-month spray period, only 247 (1.2%) constituted exposure-related health complaints. Thirty-six percent of the latter were seasonal allergy-like and flu-like symptom reports, while 17% were respiratory complaints. Interestingly, even the low level of health complaints decreased markedly as the spray program progressed, suggesting that early concerns resulting from pre-spraying publicity were allayed and that many of the reported symptoms were unrelated to B. thuringiensis exposure. Ground sprayers were exposed to 0.6–15.8  106 spores/m3 of sampled air (with cumulative exposures as high as 7.2  108 spores), while hose operators were exposed to 0.2– 8.3  106 spores/m3. These workers reported individual adverse symptoms at two or three times the control rate, with total symptom reporting rates of 63% among workers and 38% among controls. The most common symptoms were itchy skin/chapped lips, ocular irritation, dry/sore throat, runny nose/stuffy sinus, and respiratory complaints (cough, tightness). Medium exposure levels (1–3  108 total bacteria) and high exposure levels (3  108 total bacteria) registered 1.7-fold and 2.7-fold more complaints than low exposure levels (0–1  108 total bacteria), respectively. It was not clear if the symptoms were due to bacterial exposure or to other components in the spray formulation. In any case, none of the reported conditions resulted in serious illness; the number of days lost to work during the spray period was 0.16 among controls and 0.14 among workers. Interestingly, in view of the relative persistence of this bacterium in laboratory rodents, nose swabs of exposed workers were culture positive in the days following exposure; many highly exposed workers remained positive for 14–30 days, and some were still positive by the last measurement at 63 days. Despite the individual differences in persistence (which was largely a function of bacterial load, since longer nasal residence was associated with higher exposures), symptoms were indistinguishable between those that were culture negative before 30 days and those that remained positive after that time. This indicated that the acute exposure, not the prolonged carriage, resulted in the fairly minor

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health complaints that were observed. Culture positivity among patients of the local family practice physicians enrolled in the study was also strongly correlated with each of the four aerial spraying episodes, though there was no correlation with symptom reports. There were no significant differences in symptoms reported in emergency rooms or in number of emergency room visits when aerial spray days and nonspray days were compared. Two conclusions are justified from this extensive study: (1) The level of “unwellness” in the general public was not affected by the spray program; and (2) minor irritative symptoms were detected among ground spray workers (who sustained far higher exposures than the general public), though serious illness did not evolve. 2. Petrie et al. (2003) conducted a health survey among 181 people in an area of New Zealand subjected to aerial spraying with a product containing spores of B. thuringiensis subsp. kurstaki. Approximately 70% of adults and 50% of their children reported no change in health status during the period in question, though there were statistically significant increases in specific symptoms—irritations of the upper airways, gastrointestinal problems, and “neuropsychiatric” reports (e.g., sleep problems)—in smaller fractions of the population. Hay fever sufferers were more likely to report symptoms. The authors recognized the inherent problems associated with self-reporting in a study of this nature, which may have overemphasized respondents inclined toward positive reports. In addition, environmental exposures unrelated to B. thuringiensis may have had a role in the final tallies. Nonetheless, the authors were unwilling to rule out a relationship between B. thuringiensis exposure and the irritative and gastrointestinal complaints. Since 1992, the California Department of Pesticide Regulation, through its Pesticide Illness Surveillance Program, has catalogued the state’s case reports for illnesses and/or injuries in which there was at least a possibility of B. thuringiensis causation (DPR, 2008b). One hundred and eighty of the 193 reports for the 1992–2006 period in this largely occupational database did not distinguish between B. thuringiensis and several other pesticides that were concurrently applied, minimizing the usefulness of those cases in delineating trends in B. thuringiensis toxicity. The remaining 13 cases were ones in which only B. thuringiensis was applied, leaving little doubt that the reported symptoms were caused either by this active ingredient or by some other component of the formulation. With the possible exception of one case of hives (an allergenic response, considered only a “possible” result of B. thuringiensis exposure, meaning that it was not clear that B. thuringiensis or some other undefined factor was responsible), all of the latter cases were irritational in nature, with 1 case resulting in 5 days lost from work,

Chapter | 13  Microbial Pest Control Agents: Use Patterns, Registration Requirements, and Mammalian Toxicity

2 cases resulting in 2 days lost from work, and 2 cases resulting in 1 day lost from work. While it is recognized that systemic symptoms were listed among the other 180 cases, the presence of additional pesticides complicated interpretation and left in question the role of B. thuringiensis subspecies. In any event, the apparent lack of serious effects in the California database combined with the low numbers of B. thuringiensis-related cases were generally supportive of the conclusions from the epidemiologic studies discussed above. Exposure of immunocompromised individuals to B. thuringiensis was examined in an epidemiologic study conducted in an area of Oregon that had undergone spraying with B. thuringiensis subsp. kurstaki (Green et al., 1990). Bacterial cultures from patients undergoing routine exams revealed 55 that were B. thuringiensis-positive. B. thuringiensis was ruled out as a specific pathogen in 52 of those patients, leaving three cases. While these may have been opportunistic, having occurred in people with established medical conditions, exacerbation of pre-existing symptoms or development of new symptoms as a result of B. thuringiensis exposure was not ruled out. Siegel and Shadduck (1990a, 1992) discussed at length their operational definition of infectivity, which requires demonstration of multiplication of the microorganism within the host and disruption of functional or structural homeostasis. This contrasts with their definition of persistence, in which a microorganism is present in either a multiplying or quiescent state but does not disrupt the host. B. thuringiensis shows a consistent, though in some cases prolonged, clearance pattern (McClintock et al., 1995). In one study using B. thuringiensis subsp. israelensis, splenic bacterial counts following intraperitoneal injection into mice showed no decline even after 80 days (Siegel and Shadduck, 1990a). This was interpreted as evidence that multiplication had occurred within the host, though there was no sign of toxicity. Combined with the record of generally safe use, persistence of B. thuringiensis in animals does not appear to translate into human toxicity.

13.3.1.2  Bacillus cereus B. cereus is a gram-positive, catalase-positive, rod-shaped saprophyte that is, as noted previously, conspecific with B. thuringiensis. B. cereus does not produce the parasporal inclusion bodies so important to the entomopathogenic activity of B. thuringiensis. One B. cereus strain, UW85, is currently registered in the United States as a plant growth regulator and may function as a biocontrol agent against pathogenic fungi in some crop systems (Smith et al., 1993). B. cereus is medically important not only because it is an epidemiologically significant food-based pathogen but also because it is implicated in pathologies of the lung, ear, eye, gall bladder, and urinary tract, as well as being an ­ opportunistic invader in trauma or disease cases

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(Drobniewski, 1994; Goepfert et al., 1972). It was also reported to be an opportunistic pathogen in several cancer patients (Banerjee et al., 1988). Two food poisoning syndromes, a diarrheal type, associated with consumption of a diversity of food types, and an emetic type, most commonly associated with rice and pasta consumption, are caused by B. cereus (Drobniewski, 1993; Logan and Turnbull, 1999). A heat-labile enterotoxin complex consisting of two or three protein components with molecular weights of 38–57 kDa appears to be responsible for the diarrheal symptoms. A heat-stable 10-kDa peptide is associated with the emetic symptoms. Serotyping has identified the B. cereus strains most likely to produce the two syndromes. Other factors, notably hemolysins and phospholipases, are involved in the establishment of local and systemic infections (Drobniewski, 1993). In view of the medical and epidemiologic significance of B. cereus, vigilance with respect to the dissemination and application of pesticidal products containing this organism is warranted, as well as strict attention to proper strain identification. Demonstration of an inability to produce diarrheal or emetic toxins may also be useful in moderating potential risks posed by proposed commercial uses of B. cereus strains.

13.3.1.3  Bacillus sphaericus Several crystal-producing strains of B. sphaericus, an aerobic sporulating bacterium similar to B. thuringiensis, have shown promise as mosquito larvicides (Saik et al., 1990), making them potentially useful for controlling tropical diseases like malaria and filariasis (Murthy, 1997). As noted in reviews by Charles et al. (1996) and Khetan (2001), B. sphaericus produces two categories of toxin. The first category is the crystal-localized “binary” toxin, which accumulates during sporulation and consists of two proteinaceous protoxins of 42  and 51 kDa molecular weight. The 42-kDa protein, once processed in the larval midgut, is the main effector for midgut disruption. It may also have a role in determining target species specificity. The 51-kDa protein is not toxic in itself but may enhance the toxicity of the 42-kDa protein by binding to midgut cellular receptors, thus promoting binding of the 42-kDa protein. The second toxin category includes the noncrystal-localized mosquitocidal toxins – Mtx 1, Mtx 2, and Mtx 3 – 100-, 30.8-, and 35-kDa proteins synthesized during vegetative growth. Mtx 1 has ADP­ribosylating activity similar to several other bacterial toxins (pertussis toxin, cholera toxin, and Escherichia coli heatlabile enterotoxins), with which it shares sequence homologies in the catalytic domains. Schirmer et al. (2002) have demonstrated Mtx 1-mediated cytotoxicity in mammalian cell culture that is dependent on this enzymatic activity. Toxic mechanisms for Mtx 2 and Mtx 3 are undefined. However, B. sphaericus has provoked human health concerns, having been implicated in several cases of

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­ eningitis (Allen and Wilkinson, 1969; Siegel and m Shadduck, 1990b), in the formation of a lung pseudotumor in a female with a history of respiratory ill health (Isaacson et al., 1976), and as an opportunistic pathogen in a cancer patient (Banerjee et al., 1988). Interestingly, B. sphaericus isolates from the meningitis patients and from the patient with the pseudotumor did not prove pathogenic in rabbits and mice exposed intraperitoneally or intravenously, or in mice after intracerebral exposure (Siegel and Shadduck, 1990b), raising questions about the applicability of animal testing to human pathology in this case. However, it was not clear if the strains of B. sphaericus involved in these cases were relevant to current larvicidal strains. It is worth noting that B. sphaericus is a species complex, divisible into six groups based on DNA homology, with larvicidal strains clustering in homology group IIA (Krych et al., 1980). Even among larvicidal strains, the particular B. sphaericus strain designation, exposure route, and host animal strain continue to be relevant when analyzing the toxicity of these organisms in mammalian systems. A more detailed examination of two particular studies may help to illustrate this point. Shadduck et al. (1980) investigated the pathogenicity in mice, rats, and rabbits of three entomocidal B. sphaericus strains, SSII-1, 1404-9, and 1593-4, after exposure by various routes. Neither death nor systemic illness resulted under any exposure scenario. Conjunctival instillation in rabbits of up to 109 infectious units (iu) caused local lesions which were severe at the higher doses. Less severe lesions developed after instillation of autoclaved bacteria, suggesting that a portion of the pathologic response was attributable to the presence of heat-stable foreign material. Generally mild brain lesions occurred in some rats upon intracerebral injection of all three strains. In strain SSII-1 this occurred at doses as low as 1.2  106 iu/rat. Intracerebral hemorrhage was commonly noted in mice upon intracerebral injection of 3  108  iu of any of the three strains, though rabbits similarly treated with 1593-4 did not react in this way. Subcutaneous injection of strain 1404-9 resulted in an abscess in one of five mice exposed to the highest dose of 6.7  109 iu per animal. Intraperitoneal injections into rats with 3.2  108 iu of strain 1404-9 or 4.7  108 iu of strain 1593-4, respectively, were without effect. Viable bacilli were detected in eye cultures 10–14 days after conjunctival instillation into rabbits of as few as 1.2  103 iu of strain SSII-1. Persistence in the eye was also detected with injections of 109 iu of strain 1404-9 and 108 iu of strain 1593-4, though lower doses were not tested. Similarly, bacilli were detected in brain cultures 10–14 days after intracerebral injections of as few as 108 iu per rat of strain 1593-4, 1.2  108 iu of strain SSII-1, and 3.2  108 iu of strain 1404-9. Lower concentrations were not tested for strains 1404-9 or SSII-1. These results indicate that the bacterium can persist in mammals, at least under certain conditions. Whether or not bacterial replication occurred was not explored with

respect to the ocular exposure route. However, a pattern of cerebral clearance was established after intracerebral injection of 5.5  105 iu of strain 1593-4 into rats. Detections of greater than 600 iu per 100 mg wet brain tissue were noted on day 3 postinjection. No more than 30 iu/100 mg tissue were observed at various times between days 5 and 12. The brains were considered sterile by day 14. In a follow-up study, conjunctival instillation into rabbits of 4.48  108 cfu of B. sphaericus strain 2362 did not cause local toxicity (Siegel and Shadduck, 1990a), contrasting with the results of the earlier study using other strains. Nonetheless, culturable bacilli were recovered 8 weeks after treatment (longer recovery times were not examined). Splenic clearance was established after intraperitoneal injection of 1.2  107 cfu into mice, though some bacilli (165 cfu/g spleen) were still present at study termination on day 67.d Intraperitoneal injection of 8  108 cfu resulted in the death of 42 of 49 mice within 24 h; injection of half that amount (3.8  108 cfu/animal) killed all of the 6 animals treated. Interestingly, injection with autoclaved strain 2362 at 3.8  108 cfu/animal resulted in the deaths of 3/6 mice between 24 and 48 h post-­injection. These results were surprising in light of the observation of no toxicity by the intraperitoneal route of strains SSII-1, 1404-9, and 1593-4 in the earlier study (Shadduck et al., 1980). Production of a soluble extrabacterial toxin was ruled out as a cause because passage through cellulose acetate filters removed the toxicity. The authors speculated that the strain of mouse – outbred CD-1 – could have been sensitive to the effects of strain 2362, citing a study from a different laboratory showing no effect of this strain in Swiss mice. Alternatively, they considered that the particular culture conditions of the bacterium, which they did not know, could have generated a more lethal bacterial isolate. In any case, they noted that these were extremely high doses which are not likely to be relevant to human exposures in the field. Testing to support pesticidal registration of B. sphaericus strain 2362 was reviewed both by U.S. EPA and by California DPR (U.S. EPA, 1998b; DPR, 1995). Oral (3  108 cfu), intranasal (5.6  108 cfu), and intravenous (1.37  107 cfu) administrations resulted in neither mortality nor toxicity. Clearance after oral exposure was complete from all major organ systems between days 7 and 14. Intranasal exposure also resulted in complete clearance from most organs by day 14. The exception was the lung, where a clearance pattern was established by day 49

d

 The designation “colony forming units” (cfu) was used in the later study (Siegel and Shadduck, 1990a) to signify that the bacterial titers were quantitated by culturing appropriate dilutions of the inocula in agar and counting the resultant bacterial colonies at a later time. “Infectious units” (iu) was used in the earlier study (Shadduck et al., 1980) because titers were quantitated by serial tube dilution and determination of infectious units by turbidity in brain infusion broth.

Chapter | 13  Microbial Pest Control Agents: Use Patterns, Registration Requirements, and Mammalian Toxicity

(study termination). Intravenous exposure resulted in a clearance pattern from all organ systems by day 35, though bacilli remained culturable in small amounts from lungs, liver, and spleen on day 49. As implied previously, the possibility that B. sphaericus has unique pathogenic properties which are not evident in conventional animal testing (e.g., it may be infective or pathogenic only in health-compromised ­individuals) cannot yet be excluded. While there is no evidence for adverse impacts of B. sphaericus in human populations following pesticidal applications, continued attention to this possibility is warranted in view of the older medical data.

13.3.1.4  Burkholderia cepacia A gram-negative, nutritionally versatile and highly ­antibiotic-resistant bacterium, B. cepacia has generated economic interest based on its ability to inhibit soil-borne plant pathogens through the production of antimicrobial compounds, to degrade hydrocarbons associated with sites of environmental contamination, and to promote crop growth through root colonization and nitrogen fixation (Parke, 1998; Vinion-Dubiel and Goldberg, 2003). It also causes serious opportunistic infections in humans suffering from chronic granulomatous disease and cystic fibrosis (CF) (Butler et al., 1995), which may have played a role in its withdrawal by the manufacturer from the commercial market several years ago. As many as 40% of patients in some CF centers develop B. cepacia infections, with 35% of those patients exhibiting “cepacia syndrome” characterized by grave pulmonary pathogenesis, bacteremia, and death (Holmes et al., 1998). Other figures are less alarming but serious nonetheless. LiPuma (1998) cited respiratory culture results from 1996 showing that 3.6% of CF patients show evidence of infection. Of those infected, 20% “succumb to a rapidly progressive necrotizing pneumonia.” B. cepacia also is implicated in nosocomial infections of non-CF patients, as well as in the “foot rot” syndrome experienced by soldiers in swampy terrain (Holmes et al., 1998). The primary mode of transmission to CF patients appears to be from other CF patients, though transmission from non-CF patients is also possible. Social isolation measures have been necessary in some circumstances (Walters and Smith, 1993), but these have been accompanied by poor psychosocial outcomes (Butler et al., 1995; LiPuma, 1998). B. cepacia strains are currently divided among nine genomovars, which together are known as the B. cepacia complex (Vinion-Dubiel and Goldberg, 2003). According to the analysis of Vandamme et al. (1997), genomovars II and III are the best represented among CF patients. However, they caution that a systematic study of genomovar distribution has not yet been done, nor is the relative significance for cepacia syndrome of these various strains

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yet understood. Representatives of all genomovars have been detected in CF patients (Vandamme et al., 1997). Attempts have been made to identify markers associated with epidemic transmission of B. cepacia among CF patients. Such markers could simplify the risk assessment process for B. cepacia pesticidal strains, should this organism once again be considered for commercial application. Cable pili, peritrichous appendages that facilitate binding to CF mucin and airways epithelial cells (Sajjan and Forstner, 1993; Goldstein et al., 1995), may comprise one marker phenotype. These structures were identified in an epidemic strain transmitted among CF patients in clinics in Toronto and Edinburgh (Sun et al., 1995). Another candidate, a 1.4-kb DNA fragment known as the “B. ­cepacia epidemic strain marker” (BCESM) was identified in seven epidemic strains of the bacteria but was not present in nonepidemic strains (Mahenthiralingam et al., 1997). Unfortunately, neither the presence of BCESM nor the presence of cable pili is considered a certain indicator of transmission potential. The cable pilus gene was detected in only one of the seven epidemic strains examined in the Mahenthiralingam study. And, as pointed out by LiPuma (1998), the presence of nonepidemic strains in the respiratory tracts of CF patients demonstrates their colonizing ability even as they lack BCESM. In addition to these markers, other virulence factors and epidemic strain-associated markers have been considered. For example, B. cepacia has pronounced immunogenic and inflammatory activity in addition to its antibiotic resistance, all of which may be mediated by a lipopolysaccharide virulence factor (Vinion-Dubiel and Goldberg, 2003). At this point, it appears that no single marker will provide absolute predictive ability for clinically important B. cepacia strains. However, they may provide some initial screening capability when considering potential pesticidal strains. George et al. (1991) investigated the effect on CD-1 mice of intranasal instillation of 5.3  108 cfu B. cepacia strain AC1100. Ruffled fur, weight loss, and inactivity were noted during the first 2 days following treatment, with recovery evident thereafter. Increased lung weights, apparent between days 2 and 14 (study termination), were attributed to macrophage influx and endotoxin-mediated edema. Declining numbers of B. cepacia were evident through day 7 in the lungs and through day 2 in the nasal cavity. B. cepacia was also present in the gastrointestinal tract for the first 2 days after treatment. This was attributed to mucociliary evacuation from the lung to the mouth and thence to the stomach, with bacterial survival afforded by the mucus coating acquired in the respiratory system. In a later study, endotoxin-resistant C3H/HeJ mice were subjected to intranasal instillation with B. cepacia strain ATCC 25416 (George et al., 1999). Lethality was observed at as low as 2.2  108 cfu/mouse, with an approximate LD50 of 7  108 cfu/mouse. While this appears inconsistent with the relative lack of effect in mice seen at a similar dose in the

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1991 study, the considerable splay evident in the mortality data should be recognized, as well as the fact that different bacterial and mouse strains were used. Other differences from the earlier study were evident. In the 1999 study, no changes in lung weights were detected through 14 days post-treatment, possibly reflecting the endotoxin-resistant status of the mouse strain used in that study or the lower bacterial dose applied (107 cfu/mouse vs. 5.3  108 cfu/ mouse in the earlier study). Also in contrast to the 1991 study, B. cepacia appeared to be stably established in the lungs, small intestine, cecum, and large intestine through study termination on day 14. Moreover, B. cepacia was cultured from the liver and spleen at 3 h and from the mesenteric lymph nodes through 10 days. These data imply that under some circumstances, B. cepacia might gain a more tenacious hold in a mammalian system. Laboratory testing of B. cepacia isolate M36, submitted for registration in the 1990s as an active ingredient to the State of California, indicated that clearance had occurred through the feces within 7 days of oral dosing of the rat with 2.85  108 cfu. However, severe fibrous adhesions between pleural surfaces of the thoracic cavity, lungs, and pericardial sac were noted in one-third of males and gray lung coloring in two-thirds of females (DPR, 1994a). Intratracheal dosing of 1.9  108 cfu/rat resulted in pulmonary clearance by day 22, with lung discoloration evident through day 8 (DPR, 1994b). Infectivity of this strain by all routes of exposure was discounted because there was no evidence of multiplication. Significantly, use of M36 had been eliminated prior to wholesale market withdrawal due to the presence of BCESM. Because of its large and highly adaptable genome, there is also concern that pathogenic strains of B. cepacia could be generated through gene transfer or recombination if large numbers of putative nonpathogenic organisms are artificially introduced into the environment (Holmes et al., 1998). For this reason, CF advocacy groups have expressed serious misgivings about the registration of B. cepacia products before adequate testing and assurance of nontransformability is available (PTCN, 1997). As noted previously, no B. cepacia products are currently registered in the United States.

13.3.2  Eukaryotic Pesticides (Fungi and Nonfungi (Stramenopila)) 13.3.2.1  Metarhizium anisopliae In use since the late 1800s, M. anisopliae is a Deuteromycete (fungi imperfecta) employed in the United States largely for the control of cockroaches, though it is also effective against other orthopterans and against coleopterans. It has a wide geographic distribution, existing as an insect or nematode parasite or in various soils, sediments, spoil heaps, and other environments (Domsch et al., 1980).

Hayes’ Handbook of Pesticide Toxicology

Death of the host insect results when contact with conidia (the environmentally stable asexual spore stage) leads to infection. This is followed by enzyme-mediated exoskeletal degradation, mycelial development, and sporulation (Ward et al., 1998). Massing of conidia in affected insects lends a characteristic green color, hence the name “green muscardine” for the insect disease (Ferron, 1981). Insecticidal activity also may reside in a family of cyclodepsipeptides known as destruxins, 15 of which were identified in M. anisopliae as of 1989 (Gupta et al., 1989), as well as in other toxic substances (Domsch et al., 1980). Standardized laboratory studies have not demonstrated toxicity or infectivity in rats, mice, or rabbits, though persistence without multiplication was reported (Siegel and Shadduck, 1990b). CD-1 mice subjected to acute gavage with 108 conidia from M. anisopliae var. anisopliae (EH 479/2) showed neither illness nor inflammation during a 21-day posttreatment period, suggesting that M. anisopliae is nonimmunogenic and failed to germinate (Toriello et al., 2006). One mortality among the 72 animals treated was attributed to a disseminated blood coagulation event or to a state of natural immunosuppression in that individual. Allergenicity may result from M. anisopliae exposure, as demonstrated in a series of studies conducted by Marsha Ward and colleagues in mice. Allergenic responses followed intraperitoneal sensitization and intratracheal challenge with crude protein extracts of mycelia and conidia (Ward et al., 1998). Further investigation in mice showed that sensitization and subsequent intratracheal challenge with M. anisopliae crude antigen resulted in pulmonary inflammation and increased metacholine sensitivity; hyperresponsiveness was still present at 8 days post challenge. Such responses were “consistent with elevated levels of serum and BALF (bronchial alveolar lavage fluid) total IgE, BALF IL-4 (interleukin 4), eosinophils, and lymphocytes” (Ward et al., 2000). Instanes et al. (2006) showed that mice primed with a combination of the allergen ovalbumin and mycelia from M. anisopliae generated higher levels of anti-ovalbumin IgE and IgG1 when challenged 20 days later by ovalbumin than mice primed with either immunogen alone. Priming with mycelia also increased the weight and cellularity of the popliteal lymph nodes after ovalbumin challenge. Demonstration of such adjuvant effects has clinical implications since, as noted, M. anisopliae is commonly used against cockroaches, themselves potent allergens. While applicability of the mouse allergy model to humans exposed to M. anisopliae by the inhalation route has not been experimentally addressed, evidence for allergy under occupational situations has been reported (Kaufman and Bellas, 1996). A single case of keratomycosis, an increasing problem with fungi in general due to increased use of antibacterial drugs, immunosuppressants, and corticosteroids (Ishibashi et al., 1986), was reported in an 18-year-old man (Cepero de Garcia et al., 1997).

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A single case of hyphomycotic rhinitis was reported in a cat (Muir et al., 1998). Disseminated infection with severe morbidity has also been reported in an immunocompromised child (Burgner et al., 1998), highlighting the need for great caution where exposures of sick individuals to this organism are possible.

13.3.2.2  Beauveria bassiana B. bassiana, a Deuteromycete long known for its entomopathogenic properties, causes an insect disease known as white muscardine. The organism produces a number of cyclodepsipeptides such as beauvericin which may account for at least part of its toxicity to insects (Miller et al., 1983). Beauvericin may also have antimicrobial, cytotoxic, and apoptotic activity (Klaric and Pepeljnjak, 2005). B. bassiana has been used as a medicant in Japan for over a millenium (Ignoffo, 1973). Allergic responses have been reported in humans following inhalation of spore preparations, though repeated handling of cultures did not reveal adverse effects in another study (Ignoffo, 1973). A mouse study from China noted hypersensitivity-like pulmonary reactions in mice and rats after a single exposure to B. bassiana. However, the low room temperatures may have constituted a significant stress to the animals (Song et al., 1989, cited in Semalulu et al., 1992). Russian investigators determined the LD50 to be greater than 1.1  1010 and greater than 2.2  1010 fungal cells in albino rats exposed intragastrically and intraperitoneally, respectively, and greater than 4  1010 fungal cells in rabbits exposed intravenously (Mel’nikova and Murza, 1980). No significant toxicity or pathogenicity by the oral, dermal, or pulmonary routes were noted by the U.S. EPA in reviews of a series of acute studies on B. bassiana strain HF23 submitted to support its registration as a microbial pesticide, though there was mild eye irritation (U.S. EPA, 2006). Acute inhalation, hypersensitivity, and immune response studies were waived for this strain based on the evidence for clearance and the absence of toxicity in the other studies, as well as the low toxicity potential of the inert ingredients. B. bassiana has been implicated in at least two cases of keratomycosis, though both patients had long histories of antibiotic and corticosteroid use (Ishibashi et al., 1986). Separate clinical studies identified Beauveria subspecies colonizing the liver (Henke et al., 2002) or deep skin structures (Tucker et al., 2004) of immunocompromised patients under treatment for leukemia. Direct inoculation into rabbit corneas of B. bassiana isolated from a patient with keratitis resulted in inflammation, corneal ulcers, corneal haze, injection of the iris, and sparse-to-moderate fungal growth in the cornea, though the severity was less than that seen in parallel eyes treated with Candida albicans and tended to resolve with time (Ishibashi et al., 1986). Injection of B. bassiana into the quadriceps muscles of CD-1 mice led to focal muscle necrosis, edema, and inflammation, with

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the severity of the responses dependent on the number of organisms injected (Semalulu et al., 1992). Muscle regeneration was visible by 7 days. Viable spores capable of initiating colonies in artificial media were not detected after 3 days. While it is unlikely that this organism, which does not grow well at temperatures above 32°C, can infect or colonize humans under normal circumstances, it appears that exposure to Beauveria must be avoided when sick or immunocompromised individuals are present.

13.3.2.3  Gliocladium virens Also known as Trichoderma virens, this common, soildwelling saprophyte is useful in controlling Pythium ultimum and Rhizoctonia solani, organisms that cause damping-off disease in greenhouses. Such antifungal activity may be due partially to production of gliotoxin, a relatively nonselective antibiotic that is also immunosuppressive and moderately toxic to mammals (Lumsden and Walter, 1995). Studies using G. virens isolate GL-21 submitted to the California DPR to support registration did not reveal unusual toxicity, though conjunctival irritation for at least 72 h in the rabbit eye irritation study and death by capillary obstruction in the rat intravenous study were noted (DPR, 1993). The potential for G. virens-induced allergenicity is not known, though gliotoxin can alter the secretion patterns of pro- and anti-inflammatory cytokines in human monocytic cell cultures (Johannessen et al., 2005).

13.3.2.4  Lagenidium giganteum The genus Lagenidium is a member of the class Oomycetes. Oomycetes are not true fungi but are rather members of the kingdom Stramenopila, which also includes planktonic diatoms and multicellular seaweeds. The ability of L. giganteum, an aquatic saprophytic oomycete, to parasitize, and eventually kill, mosquito larvae underscores its potential in mosquito control programs. Conventional toxicity testing with an organism as large as L. giganteum, which can produce cells greater than 200 m in length, was difficult because intratracheal instillation into rats of as few as 1.16  105 oospores resulted in the prompt death of many of the treated animals from acute pneumonia, airways obstruction, or severe inflammation (Siegel and Shadduck, 1987). A similar result was obtained upon intravenous injection of 1.78  106 cfu into mice, where embolism killed several animals within 36 h of treatment (Kerwin et al., 1990). Lowering the numbers of active or autoclave-inactivated organisms still resulted in tissue damage after intratracheal or intraperitoneal exposures (Siegel and Shadduck, 1987). Such lesions may represent local inflammatory responses to large amounts of foreign biological material. While there was histologic evidence for persistence, multiplication within the mammalian

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hosts did not occur. Oospore treatment of rat skin or rabbit eyes did not result in irritation. The potential for allergenicity remained untested. Grooters et al. (2003) identified a Lagenidium species as the causative agent in six dogs with devastating infections characterized by cutaneous lesions, lymphadenopathies, and related conditions. Serologic and molecular analysis indicated similarities between the infective agent and L. giganteum, though the variant growth and morphologic characteristics, as well as the inability of L. giganteum to infect rodents, suggested a distinct Lagenidium species in the dog cases (personal communication, A. Grooters, Louisiana State University; Grooters et al., 2003). Nonetheless, the fact that a closely related Lagenidium species can be so pathogenic to dogs should serve as a reminder that current understanding of the potential for human infection and pathogenicity mediated by L. giganteum may be inadequate.

13.3.3  Viruses Interest in the use of viruses as pest control agents has increased in recent years due to their promise as specific vectors, either in their native form or as genetically engineered constructs designed as delivery agents for pesticides or as immunocontraceptives for mammalian wildlife control. Establishment of the range of target species susceptible to infection by a given virus is perhaps the major issue in the assessment of risks associated with use of that virus as a pest control agent. For obvious reasons, human health concerns are magnified when the virus in question is infective in, or pathogenic toward, mammalian species.

13.3.3.1  Baculoviruses Baculoviruses are double-stranded DNA viruses. Two major groups of baculoviruses, the nucleopolyhedrosis viruses (NPVs) and the granuloviruses (GVs), are enveloped by proteinaceous occlusion bodies that protect the virions from adverse environmental impacts (Saik, 1990). Infection in susceptible insects occurs when the occlusion bodies are ingested from leaf surfaces and dissolved under the alkaline conditions of the insect gut (Black et al., 1997). Death of the host results from the multiplication of freed virions within bodily tissues (Huber, 1995). GVs are of less practical use than NPVs for insect control, as they are more difficult to grow in cell culture (Khetan, 2001). Based on studies both in humans and in laboratory animals, human safety concerns appear to be minimal. Dietary consumption over a 5-day period of 6  109 Helicoverpa zea NPV polyhedra was wi444thout effect (Heimpel and Buchanan, 1967). Occupational exposure to Baculovirus heliothis was similarly without effect over a 26-month period in a virus production facility (Huang et al., 1977). Exposure of mice and guinea pigs to H. zea

Hayes’ Handbook of Pesticide Toxicology

NPV inclusion bodies, virus rods, and polyhedral protein by several routes (inhalation, oral gavage, and intradermal, intraperitoneal, intracerebral, and intravenous injection) revealed no effects (Ignoffo and Heimpel, 1965). Similar negative results were obtained using the Trichoplusia ni NPV (Heimpel, 1966). No unusual responses were detected in rats following acute subcutaneous injection of 1.2  109 polyhedral inclusion bodies (PIB) of H. zea NPV into neonates or acute intravenous injection of the same quantity into adults (Barnes et al., 1970). In addition, no effects were identified with dietary exposures of 90 days or 2 years using feed preparations containing viral loads between 6  107 and 6  109 PIB/100 g. Allergenic responses were not detected in guinea pigs after 3 weeks of inhalation exposure at 1 h/day, 5 days/week, to H. zea NPV free viral rods (obtained from 3  1011 PIB/day) or intact PIB (3  1011/day), or after dermal exposure to 1.5  1011 PIB for 5 days followed by intradermal injections of 1.2  108 PIB in each of four bodily sites (Meinecke et al., 1970) Pigs force-fed with Mamestra brassicae larval NPV at a dose level of 5  107 polyhedra/g body weight showed a slight, transitory rise in body temperature (Doller et al., 1983). There was, however, no evidence for lymphatic involvement, no effect on leukocyte counts, and no evidence for viral replication or organ infection within the hosts. In one account published in 1965, over 26 baculoviruses had been tested in 10 mammalian species without indication of toxicity (Ignoffo and Heimpel, 1965). Indeed, baculoviruses do not appear capable of multiplying within mammalian hosts (Black et al., 1997). Finally, the ubiquity of baculoviruses in the human food supply attests to their harmlessness (Black et al., 1997; Heimpel et al., 1973).

13.3.3.2  Rabbit Hemorrhagic Disease Virus Rabbit hemorrhagic disease virus (RHDV), a singlestranded RNA virus belonging to the calicivirus family, is used as a rabbit biocontrol agent in Australia and New Zealand. It is linked to a syndrome of necrotizing hepatitis, hemorrhage, and death in European rabbits (Oryctolagus cuniculus), known as rabbit hemorrhagic disease (RHD). RHD was first recognized in Angora rabbits exported from Germany to China in 1984, spreading thereafter to Korea, Europe (including the British Isles), Mexico, Israel, and North Africa (Nowotny et al., 1998). RHDV escaped from quarantine and was then purposefully dispersed in Australia, where introduction of European rabbits in the 19th century left the island continent with a staggering rabbit infestation. The rabbits’ prodigious appetite, burrowing practices, and reproductive capacity have wreaked ecological and agricultural devastation, with serious consequences for the Australian economy. The propagation of RHDV/ RHD, whether by deliberate or natural means, has reduced European rabbit populations in recent years (Cooke,

Chapter | 13  Microbial Pest Control Agents: Use Patterns, Registration Requirements, and Mammalian Toxicity

2002), though viral efficacy can vary depending on such host and nonhost factors as rabbit age, density, resistance, presence or absence of vectors, and local climatic conditions; indeed, such factors underlie regional heterogeneity in disease impacts (Cooke, 2002; Parkes et al., 1999). It is nonetheless undeniable that RHDV puts rabbit populations at risk regardless of whether or not eradication is a desired endpoint. Like Australia, New Zealand has considered using RHDV to control its own rabbit infestation, though importation was initially prohibited in 1997 due to uncertainties surrounding the epidemiology and efficacy of the virus (Ministry of Agriculture and Forestry, 1997). Despite this ruling, the virus became established in New Zealand by virtue of illegal importation and release (Sissons and Grieve, 1998). Deliberate spread by farmers, while initially of unclear legality, eventually led to legalized use through an amendment to the law, provided it is done with appropriate permits (personal communication, J. Parkes, Landcare Research, New Zealand). Following government-sanctioned tests of the efficacy of rabbit-to-rabbit transmission conducted in 1995 on Wardang Island, an uninhabited tract several kilometers off the coast of southern Australia, RHDV jumped to the mainland by an unknown mechanism. This led to the approval of hundreds of deliberate releases on the mainland and brought into focus the issue of human safety in the affected areas. The wide host ranges of other calicivirus types, along with the virulence of some caliciviruses in humans [representatives of four of five calicivirus categories are considered to be human pathogens (Smith et al., 1998)] and the high mutation rates characteristic of RNA viruses in general, caution against premature judgments on the issue of possible direct human impacts. The establishment of credible evidence for RHDV specificity toward European rabbits is considered an important step in ensuring its safety as a pesticide. Current approaches to this question are based on antibody analysis of blood from a variety of experimentally infected ­ species (Bureau of Resource Sciences, 1996), as well as on tests to detect viral RNA replication in species that have demonstrated seropositivity (Gould et al., 1997). Despite the lack of evidence for overt disease in any species outside of European rabbits, several studies have shown seroconversion after inoculation with, or dietary exposure to, RHDV (Buddle et al., 1997; Leighton et al., 1995; Zheng et al., 2003). In addition, Parkes et al. (2004) demonstrated seroconversion in several wild vertebrate species in New Zealand, including feral cats (38/71), stoats (2/8), ferrets (11/115), hedgehogs (4/73), hares (3/66), hawks (2/18), and gulls (1/30). These are, for the most part, rabbit predators or scavengers, characteristics that likely account for the exposure route. The data in hares, on the other hand, may reflect infection with a cross-reacting hare-specific calicivirus or the presence of an unknown vector, as hares are not known to consume rabbits. Oral and/or anal excretions from flies have tested

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positive in areas of Australia where the disease is endemic (Asgari et al., 1998; Cooke, 2002), suggesting a means of transmission independent of ingestion of diseased rabbits. Insect-mediated transmission probably also explains the island-to-mainland jump alluded to previously. In any event, when dealing with serologic data, rigorous tests are required to distinguish active infection from simple exposure. Such a distinction is important because it bears on the question of the ease by which the virus crosses species barriers as an active infective agent. Gould et al. (1997) examined this question directly in 28 experimentally infected nonhost species using a polymerase chain reaction to detect viral RNA; in no case was infection detected despite a species range spanning domestic, feral, and wildlife animals. Along with the lack of evidence for overt disease outside of European rabbits, Gould’s data suggest that the possibility of nonrabbit viral replication and pathogenicity is low. With this background, it is interesting to note that where the possibility of human infectivity or illness resulting from contact with infected rabbits has been examined directly, divergent conclusions have been drawn, even when they are based on the same data. Mead et al. (1996), in a report to the Australian government, as well as Carman et al. (1998) in the follow-up report in the open literature, found no evidence for infection or for symptoms of disease in over 250 people, many of whom had high exposure to the virus through direct handling of diseased rabbits. Smith et al. (1998), using the same data but categorizing putative exposure levels differently, argued for correlations between exposure and incidence of a number of pathologies, including flu/fever, diarrhea/gastroenteritis, neurologic symptoms, rashes, and hepatitis. Matson (1998) also differed from Mead in his analysis of the human serologic data provided in the original Mead report, seeing plausible evidence for infection in some people from South Australia who had handled infected rabbits. However, Greenslade et al. (2001) saw no evidence of seroconversion in separate cohorts of heavily and moderately exposed people from two rural communities in New Zealand. Because of these conflicting data and interpretations, clarification of the issue of RHDV specificity and safety to humans will await rigorous analysis of well-planned future studies of both an epidemiologic and an experimental nature.

Conclusion The promise of microbial pest control agents resides in their host organism specificity and in their relatively benign ecosystem and human health impacts. The requirements for toxicity testing of these agents in the United States under 158 CFR 40 are designed to provide a fast and efficacious means to identify problematic MPCAs, while ­ moving the rest toward registration. In general, this approach appears to have functioned well. Nonetheless, as is recognized

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in those ­ guidelines, MPCAs pose unique challenges to humans. They are, after all, alive, and thus carry at least a theoretical potential for adaptation and survival in novel microenvironments. It is possible that hazards to humans posed by living organisms could be missed in animal studies. Particular attention must be paid to the welfare of sensitive human subpopulations such as those who are diseased or immunocompromised, or who might be allergic to specific microorganisms. Attention to strain type is always important, and even more so when the microbe in question belongs to a medically significant species or genus. In the final analysis, continued monitoring of the health effects of MPCAs under conditions of actual pesticidal manufacture and use is warranted to ensure the safety of this fascinating and viable approach to pest control.

Acknowledgments I acknowledge the following: for helpful discussions and comments on the manuscript (in whole or in part), Amy Grooters, Svetlana Koshlukova, John Parkes, and Alan Reynolds; for the transmission electron micrograph of Bacillus thuringiensis and its accompanying caption, Brian Federici; for help accessing and interpreting data from the Pesticide Use Report (California Environmental Protection Agency), Larry Wilhoit; for help accessing and interpreting data from the Pesticide Illness Surveillance Program (California Environmental Protection Agency), Louise Mehler.

References Allen, B. T., and Wilkinson, H. A. III (1969). A case of meningitis and generalized Shwartzman reaction caused by Bacillus sphaericus. Hopkins Med. J. 125, 8–13. Angus, T. (1954). A bacterial toxin paralyzing silkworm larvae. Nature 173, 545–546. Asgari, S., Hardy, J. R. E., Sinclair, R. G., and Cooke, B. D. (1998). Field evidence for mechanical transmission of rabbit haemorrhagic disease virus (RHDV) by flies (Diptera: Calliphoridae) among wild rabbits in Australia. Virus Res. 54, 123–132. Banerjee, C., Bustamante, C. I., Wharton, R., Talley, E., and Wade, J. C. (1988). Bacillus infections in patients with cancer. Arch. Intern. Med. 148, 1769–1774. Barnes, R. W., Meinecke, C. F., McLane, W. C., and Rehnborg, C. S. (1970). Long-term feeding and other toxicity pathogenicity studies on rats using a commercial preparation of the nuclear-polyhedrosis virus of Heliothis zea. J. Invert. Pathol. 16, 112–115. Beegle, C. C., and Yamamoto, T. (1992). Invitation paper (C.P. Alexander Fund): history of Bacillus thuringiensis berliner research and development. Canadian Entomologist 124, 587–616. Bernstein, I. L., Bernstein, J. A., Miller, M., Tierzieva, S., Bernstein, D. I., Lummus, Z., Selgrade, M. K., Doerfler, D. L., and Seligy, V. L. (1999). Immune responses in farm workers after exposure to Bacillus thuringiensis pesticides. Environ. Health Perspect. 107, 575–582.

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Chapter 14

The Assessment of the Chronic Toxicity and Carcinogenicity of Pesticides Deborah Ramsingh Pest Management Regulatory Agency, Health Canada, Ottawa, Ontario, Canada

14.1  Introduction A rigorous and thorough assessment of the chronic toxi­ city and carcinogenic potential of a pesticide is integral to the evaluation and regulation of pest control products. While pesticides are beneficial to humans in many ways, they are a group of chemicals of particular concern due to their deliberate introduction into the environment and their often inherently toxic nature that is required to exert their control over unwanted pests. Humans in the general popu­ lation may incur long-term exposure to pesticides through the ingestion of residues in/on food and in drinking water. Individuals working with pest control products (e.g., farm­ ers and exterminators) have the potential to be exposed to pesticides over an extended period through other routes of exposure (i.e., dermal and inhalation), as do residents through the use of pesticides in and around the home. For these reasons, pesticides are highly regulated and require approval by authorities prior to use. In order to ensure an appropriate level of human safety, regulating authorities must ensure that levels of pesticides to which humans may be exposed are within acceptable limits. An evaluation of the hazard posed by long-term exposure to a pesticide is conducted and the highest exposure level that is not likely to result in deleterious effects is determined. This is accomplished by gathering extensive information from chronic animal toxicity studies in which the pesticide is administered daily to test animals over a major portion of their lifespan. Lifetime exposure of the animals is an impor­ tant component of these toxicity studies since the response to a particular chemical may change with age-related altera­ tions in tissue sensitivity, metabolism, and disease states

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[World Health Organization (WHO), 1978] or as a result of prolonged exposure. Experimental animals are consid­ ered to be well-understood predictors of toxicity in humans and thus useful models for the detection of potential human responses to pesticides (Gad et al., 2007; Haggerty et al., 2007; Johnson and Gad, 2007). The majority of chemicals known to be carcinogenic to humans have tested positive for carcinogenicity in animal studies (Jacobs, 2006). The purpose of chronic toxicity studies is multifaceted: (1) to identify target organs of toxicity; (2) to describe dose–response relationships; (3) to determine pos­ sible effects of cumulative exposure; (4) to generate data on the development of pre-cancerous lesions; (5) to permit assessment of the potential for tumor development; (6) to provide insight into the mode or mechanism of action; and (7) to inform the analysis of the weight of the evidence when integrating results from other toxicity studies. Extensive information on systemic toxicity, including physiological, hematological, and biochemical effects, as well as carcinogenic potential, is generated through the observation of the animal’s behavior, the examination of organs and tissues, and the clinical analysis of blood and urine. A variety of species are used in these studies to indi­ cate whether the same effects are observed in different spe­ cies or if effects are limited to a certain species. Studies for chronic toxicity testing are usually designed in such a way that the lowest dose that results in adverse eff­ ects, or the lowest-observed-adverse-effect level (LOAEL), and the highest dose that does not elicit adverse effects, or the no-observed-adverse-effect level (NOAEL), can be estab­ lished for noncancer endpoints, which are generally assumed to act through a threshold process (i.e., doses low enough

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will not elicit any observable toxicity). By contrast, cancer is most often considered a nonthreshold effect (i.e., effects are assumed to occur at any exposure level) in the absence of data demonstrating otherwise. Thus, the primary aim of car­ cinogenicity studies is to determine the presence or absence of a carcinogenic response and the potency of that response. The information obtained through the assessment of chronic toxicity and carcinogenicity studies is integrated with results from other toxicity studies conducted with the pesticide, as well as information on structure–activity rela­ tionships, mode of action information, and epidemiology data. This collective knowledge is used to determine the exposure levels considered to be acceptable for humans. The potential risks from anticipated exposures to the pes­ ticide are then estimated by comparing these acceptable levels to the amount of pesticide to which humans may be exposed over the long term. The intent of this chapter is to provide an in-depth description of the process through which the chronic haz­ ards and carcinogenic potential of pesticides are assessed. The internationally accepted protocols normally used for the chronic toxicity and carcinogenicity testing of pes­ ticides are outlined. The standard endpoints assessed in these protocols and how these endpoints are interpreted and translated into manifestations of toxicity are presented. Approaches to carcinogen hazard assessment are dis­ cussed. Only a brief outline of the way in which chronic toxicity and carcinogenicity information from animal toxi­ city studies is applied to the risk assessment is provided, as approaches to risk assessment are jurisdiction-specific and somewhat varied. Finally, some of the limitations of the current testing paradigm are discussed. As there is a diverse array of toxic effects following chronic exposure, it is impossible to summarize in this chapter known toxici­ ties following long-term exposure to various pesticides. A publication summarizes the chronic toxicity profiles of 310 pesticides reviewed for registration in the United States (Martin et al., 2009). A compilation of the carcinogenic classification of 465 pesticides has been released by the U.S. Environmental Protection Agency (EPA, 2006).

14.2  Regulatory requirements, test guidelines, and protocols Several pesticide regulatory bodies, such as those in the United States, Canada, the European Union, Australia, New Zealand, and Japan, have similar requirements for chronic toxicity and carcinogenicity testing of pesti­ cides. These requirements are based largely on test guide­ lines published by the U.S. EPA and the Organization for Economic Co-operation and Development (OECD). These guidelines describe how animal bioassays should be conducted for the assessment of chronic toxicity and carcinogenicity.

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Separate guidelines have been prepared for chronic tox­ icity testing (OECD, 1981a, 2009a; U.S. EPA, 1998a) and carcinogenicity testing (OECD, 1981b, 2009b; U.S. EPA, 1998b), while another guideline outlines a study protocol that allows for the testing of noncancer and cancer end­ points to be combined (OECD, 1981c, 2009c; U.S. EPA, 1998c). A summary of the study protocols recommended in these guidelines is outlined in Table 14.1. To be consid­ ered acceptable for regulatory purposes, studies must be conducted according to the principles of good laboratory practice (Health Canada, 1998; OECD, 1998; U.S. EPA, 1989). While the majority of chronic toxicity and carcino­ genicity studies conducted with pesticides adhere for the most part to recommendations put forth in these guide­ lines, deviations may be considered acceptable by regula­ tory authorities on a case-by-case basis, when warranted and properly justified. Above all, toxicity studies should be designed to be “scientifically meaningful” (WHO, 1978), regardless of regulatory requirements.

14.2.1  Species and Study Duration For the evaluation of pesticides, chronic toxicity testing is normally required with two mammalian species, one a rodent and the other a nonrodent, to account for dif­ ferences in pharmacokinetics and pharmacodynamics between test species and humans (Gad et al., 2007). The rat is the preferred rodent species while the dog is the pre­ ferred nonrodent species. The rat’s practical size, meta­ bolic similarities to humans, relatively docile nature, fairly constant genetic profile, and short lifespan and the dog’s size, even temperament, and relatively low cost provide advantages to using these species for toxicity testing over others (Haggerty et al., 2007; Johnson and Gad, 2007; Keller and Banks, 2006). Dogs require additional housing and exercise, thus demanding more laboratory space and additional personnel, and usually have a wide variation in size and body weight (Haggerty et al., 2007), but their size also permits the sampling of larger volumes of blood when compared to rodents (WHO, 1978). Mice are not normally used for chronic toxicity testing since some procedures may be hampered by their small size (Keller and Banks, 2006). Also, mice tend to be more sensitive to environmen­ tal deviations (e.g., failure of the watering or air-conditioning systems) than rats due to their smaller size and faster metabolism (Gad et al., 2007). Dosing for a minimum of 12 months is required for chronic toxicity studies. For the hazard assessment of pes­ ticides, most chronic toxicity studies are 12–24 months in duration, with studies in dogs typically lasting 12 months and those in rats lasting 24 months. However, some regula­ tory authorities do not consider a 1-year study in the dog to be a true chronic toxicity study as it covers less than 50% of the lifespan of this species (WHO, 1990). These

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Table 14.1  Summary of Protocols Outlined in U.S. EPA and OECD Test Guidelines for Chronic Toxicity, Carcinogenicity, and Combined Chronic Toxicity/Carcinogenicitya Parameter

Chronic

Carcinogenicity

Combined

Species

Two mammalian—rodent and nonrodent: preferably rat and dog

Two mammalian: preferably rat and mouse

Usually rat

Age

As soon as possible after weaning and acclimatization: no later than 8 weeks for rodents; between 4 and 6 months but no later than 9 months for dogs

As soon as possible after weaning and acclimatization; preferably before 8 weeks of age

Sex

Male and female

Group sizes

Rodent:  20/sex Nonrodent:  4/sex

At least 50/sex

Satellite group

10–20/sex for rodents, 4/sex for nonrodents; high-dose and control groups for reversibility, if used

Not required

Doses

At least 10/sex for evaluation of chronic toxicity at 12 months

At least three plus control

Limit dose

1000 mg/kg bw/day

Frequency of dosing

7 days/week (5 days/week acceptable if dosing by gavage, capsule, dermal, or inhalation)

Duration of dosing

At least 12 months

Clinical observations

At least 18 months for mice and 24 months for rats At least 2  /day for morbidity/mortality At least 1  /day for general clinical observations Detailed clinical examination at least 1  /week

Neurological examination

Near end of first year, not earlier than month 11, assessment of motor activity, grip strength, sensory reactivity to stimuli of different types (U.S. EPA)

Body weight

Not required

Near end of first year, not earlier than month 11, assessment of motor activity, grip strength, sensory reactivity to stimuli of different types (U.S. EPA)

Weekly during first 13 weeks, monthly thereafter b

Food/water consumption

Weekly during first 13 weeks, at 1- to 3-month intervals thereafter

Hematology, clinical chemistry, urinalysis

10/sex/group for rodents; all animals for nonrodents At 3- to 6-month intervals

Blood smear from all animals At 12 and 18 months and at terminal sacrificec

10/sex/group for rodents; all animals for nonrodents At 3- to 6-month intervals

Ophthalmoscopic examination

All animals at study start, 10/sex rodents and all nonrodents at termination; at least control and high dosed

Not required

All animals at study start, 10/sex rodents and all nonrodents at termination; at least control and high dosed

Macroscopic examination

All animals

Microscopic examination

All grossly visible tumors and other lesions; target tissues in all animals; all animals that died or were killed during the study; all animals in the control and high-dose groupsd

Organ weights a

All animals

10/sex/group (not required for OECD)

All interim sacrifice animals and 10/sex at termination

Adapted from OECD (1981a,b,c, 2009a,b,c) and U.S. EPA (1998a,b,c). Water consumption is required to be measured if the test material is administered in the drinking water. c Differential blood count determined for control and high-dose group at sacrifice; the 12- and 18-month blood smears should be examined and other dose groups as required. d Examination should be extended to all dose groups if treatment-related changes are observed in the high-dose group. b

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authorities may not require a 1-year dog study depending on results of studies of shorter duration (e.g., 90 days) in the dog. The endpoints examined in chronic toxicity stud­ ies include behavioral changes, growth, survival, and organ system toxicity. An elaboration of these endpoints is pro­ vided in Section 14.3. As with chronic toxicity testing, carcinogenicity stud­ ies are normally required to be conducted in at least two mammalian species. Rodents are the preferred species for carcinogenicity testing due to their small size and rela­ tively short lifespan and because large groups are required to detect treatment-related changes. Mice and rats are most commonly used and are exposed to the test chemical for a minimum of 18 and 24 months, respectively, with exposure being extended for up to 24 and 30 months for strains with greater longevity or a lower spontaneous tumor rate. The incidence of spontaneous tumors greatly increases beyond 18 and 24 months of age in mice and rats, respectively, lead­ ing to a high background tumor rate that can confound the interpretation of tumorigenicity. While not normally used for pesticide testing, hamsters and nonhuman primates are also acceptable species as stipulated in the test guidelines. Rabbits are not generally considered by the test guidelines to be an acceptable species for this type of testing. Although not commonly used for pesticide evaluation at this time, there are also some short-term bioassays that screen for neoplastic or pre-neoplastic endpoints and can provide information on the mechanism of carcinogenesis. Some of these assays use genetically engineered rodent models that are designed to accelerate tumor development and can provide results within 1 year or less (McGregor, 2006; OECD, 2002b). As indicated previously, the U.S. EPA and the OECD allow for the chronic toxicity and carcinogenicity assess­ ment of pesticides to be combined into one study (Health Canada, 2005; OECD, 1981c, 2009c; U.S. EPA, 1998c). Rats are usually used in this type of study, in which satel­ lite groups are sacrificed at 12 months for the evaluation of chronic toxicity endpoints while a main group of animals continues to be exposed for up to 24 months to observe tumor development. Combining the assessment of chronic toxicity and carcinogenicity into one study economizes use of the test animals (Health Canada, 1991) and is common practice in pesticide testing.

14.2.2  Route of Administration The oral route is most commonly used in the testing of pesticides that are likely to come into contact with food or water, with administration usually by dietary admixture. However, depending on several factors such as the physical and chemical characteristics of the test material, other pos­ sible routes of human exposure, a concern over instability or palatability in the diet, or the potential for local injury at

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the administration site, the chemical may need to be inves­ tigated by the dermal and/or inhalation route or through other forms of oral administration such as gavage, capsule (for nonrodents), or drinking water. Dietary administration may result in higher doses to animals early in the study when young animals are still growing and food consump­ tion is highest per unit body weight. While bolus admin­ istration through gavage or capsules may allow for the delivery of a more steady and well-defined dose compared to dietary administration, it may cause irritation of the ali­ mentary canal and/or result in unusual pharmacokinetics (OECD, 2002b). Stability and homogeneity of the pes­ ticide in the vehicle used must be demonstrated, and any potential reactions with components of the diet to produce toxic by-products (WHO, 1990) must be considered.

14.2.3  Dose Levels An important factor in study design is the selection of appropriate dose levels. Doses for chronic toxicity and car­ cinogenicity studies are normally based on results of shorter term toxicity studies or dose range-finding studies and should be appropriately spaced to detect a dose–response relationship for the effects observed. As a matter of course, a vehicle control group and three test material treated groups are used; however, a second untreated control group may be required if the toxicity of the vehicle is unknown or if the method of test material administration is expected to affect the test animals (Keller and Banks, 2006). The low dose in a chronic toxicity or carcinogenicity study should not produce any evidence of toxicity and thus form the basis for the NOAEL, while the intermediate dose level(s) should be spaced appropriately between the low and high doses to produce a continuum or progression of toxic effects. Anticipated human exposures should be taken into account when selecting the low and intermediate dose levels to adequately characterize the dose–response curve (ILSI, 1997; Rhomberg et al., 2007). Excessive doses that alter the physiology of the test animal so far from normal that the animal no longer represents an adequate model for human toxicity may be inappropriate for human health risk assessment (ICH, 2008). Much consideration should go into the selection of the highest dose in carcinogenicity studies, as tumor develop­ ment may be affected if the administered doses overload or saturate metabolic processes, cause severe tissue dam­ age, or affect body weight to a significant degree. In addi­ tion, to maintain statistical power and minimize the chance of observed false-positive or -negative trends, the selected doses must not result in excessive mortality. The OECD and U.S. EPA test guidelines for carcinogenicity studies stipu­ late that the highest dose level should be sufficiently high to elicit signs of toxicity without substantially altering the normal lifespan due to effects other than tumors (OECD,

Chapter | 14  The Assessment of the Chronic Toxicity and Carcinogenicity of Pesticides

1981b; U.S. EPA, 1998b). Test guidelines also stipulate that the concentration of test material in the diet should not exceed 5%, known as the maximum feasible dose (ICH, 2008), so as not to interfere with the nutritional quality of the diet (OECD, 1981b), and that the highest dose to be tested should not exceed 1000 mg/kg bw/day (U.S. EPA, 1998b). The recommendation not to exceed a dietary level of 1% for a pesticide has also been put forth (WHO, 1990). Although definitions and interpretations have var­ ied over time and among regulatory authorities, the term “maximum tolerated dose” (MTD) is used to describe the highest dose that does not alter the normal lifespan of the animal due to effects other than cancer and thus does not result in effects that would compromise the interpretabil­ ity of the study. Body weight is often used as a general indicator for the attainment of the MTD. Most authori­ ties agree that the highest dose in a carcinogenicity study should be based on a dose that causes decrements in body weight gain approaching 10% in subchronic studies (ILSI, 1997; Rhomberg et al., 2007) and should not result in body weight decrements of greater than 10–12% from controls in a long-term study (OECD, 2002a,b). However, these criteria should not be relied upon in isolation. A weight of evidence/integrative approach should be taken when deter­ mining if adequate doses have been achieved. For example, the highest dose may not result in effects on body weight or survival, criteria normally used to assess the attainment of an MTD, but instead may result in changes to another endpoint of concern, such as hepatocytic degeneration or necrosis (Rhomberg et al., 2007). Additional criteria for attainment of the MTD have been developed by ECETOC (1996), the U.S. EPA (2003), and Health Canada’s PMRA (as summarized in Rhomberg et al., 2007). There continues to be much debate around the issue of the MTD. It is often difficult to predict the value of the MTD for a study in which the test chemical is administered daily for the animal’s lifetime based upon results of shorter term studies (Mastorides and Maronpot, 2002). There may be a fine line between a dose that is too low and one that is too high. For instance, if a carcinogenicity study uses a dose that is considered to be too low, the adequacy of the study and its ability to detect a carcinogenic response is questioned. On the other hand, a dose that is too high that results in a positive response may not be relevant to human risk assessment (Mastorides and Maronpot, 2002). Despite the importance of noncancer endpoints to the assessment of carcinogenicity, the combination of chronic toxicity and carcinogenicity testing in the same study can complicate the selection of appropriate doses. This is due to the fact that the motive behind a chronic toxicity study (i.e., characterization of the threshold for noncancer end­ points) differs from that for a carcinogenicity study (i.e., characterization of the cancer hazard), resulting in dispa­ rate concerns regarding statistical power (ECETOC, 1996; Rhomberg et al., 2007). In cancer bioassays, the primary

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concern is the ability to detect rare tumors, whereas the concern regarding chronic toxicity endpoints, which are generally graded on a continuum from less to more severe, is the ability to distinguish an affected animal from one that shows no effect. However, there are some similarities in the factors that affect dose selection when assessing both cancer and noncancer endpoints – that is, dose–response assessment in addition to hazard characterization and mode of action determination, the use of doses that mimic human exposures as closely as possible, and adequate sample sizes to ensure statistical power (Rhomberg et al., 2007).

14.2.4  Chemical Purity Toxicity testing is normally conducted using the techni­ cal grade of the pesticidally active ingredient found in pest control products (Health Canada, 1991). The chemical purity of the sample used in testing, and how that purity compares to the commercial product, must be taken into consideration. The toxicity of a pesticide may be due to, or modified by, impurities. Toxic effects actually caused by exposure to impurities present in the test material may be incorrectly attributed to the pesticide itself or would not be observed if the sample of material used in a study is highly purified (WHO, 1978). Ideally, the test material used in toxicity testing should match as closely as possible the composition of the pesticide manufactured commercially (Keller and Banks, 2006).

14.3  Assessment of chronic toxicity (noncancer) endpoints The OECD and U.S. EPA test guidelines for chronic toxi­ city studies require a multitude of endpoints to be evalu­ ated (Table 14.1). Animals are monitored throughout the study for evidence of toxicity, such as effects on growth and development, food and water consumption, and mortality rate; blood and urine samples are collected periodically for the analysis of hematology and clinical pathology; and rep­ resentative target and nontarget organs are examined macroand microscopically at study termination (Barile, 2008). Many of these endpoints are not specific to chronic toxicity testing and are evaluated in other toxicity studies as well. The relation to treatment and the toxicological signifi­ cance of any alterations in the endpoints must be evaluated. Scientific judgment forms the basis of such an evaluation, and the considerations used in this evaluation are best described by criteria outlined by Lewis et al. (2002). To assess relation to treatment, considerations such as dose response, the presence of outliers, precision of measure­ ment, normal biological variation, and biological plausibil­ ity of an effect are taken into account. Comparison to both concurrent control values and historical controls, when

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available, is conducted. Consideration taken in deciding the adversity of an effect includes whether or not the effect (1) alters the general function of the test organism or organ/ tissue affected; (2) is an adaptive response; (3) is transient and disappears during the course of treatment; (4) is of min­ imal severity below thresholds of concern; (5) is isolated; (6) is secondary to other effects; or (7) is a consequence of the experimental model or treatment regimen (Lewis et al., 2002). In general, treatment-related effects classified as being nonadverse are those that do not affect morphol­ ogy, growth, development, or lifespan; do not compromise the animal’s ability to adapt to stress; are temporary and do not result in permanent damage; and do not increase the animal’s susceptibility to other environmental influ­ ences (WHO, 1978). Examples of effects noted following pesticide exposure that are not considered adverse include plasma cholinesterase depression and slight increases in liver weight that are shown to be reversible (WHO, 1990). Concurrent control groups should be used for compari­ son when assessing the relation to treatment of endpoints assessed in chronic toxicity studies. Control groups should be the same age as animals in treatment groups since some toxic effects may be due to an altered rate or degree of nat­ ural changes related to age. For example, lesions such as amyloidosis, chronic progressive glomerulonephropathy, peripheral nerve degeneration, and certain tumors occur naturally in aged animals, but their incidence and sever­ ity may be increased in a dose-related manner following exposure to a toxic agent (OECD, 2002a). Concurrent con­ trol groups may be compared to historical control data to ensure that the control groups are behaving as expected. The evaluation of equivocal treatment-related changes from concurrent controls can also be assisted by a compar­ ison to historical control values. An in-depth discussion regarding the interpretation of all endpoints assessed in chronic toxicity studies is beyond the scope of this chapter. The following paragraphs touch on the interpretation and toxicological significance of cer­ tain endpoints, as well as important factors to consider when assessing these endpoints.

14.3.1  Mortality Effort should be made to determine cause of death in ani­ mals that die prior to terminal sacrifice to aid in determin­ ing the toxicological significance of any apparent patterns in mortality rates. It is important to distinguish toxicityrelated deaths from deaths caused by other factors deter­ mined to be unrelated to treatment, such as intercurrent infections, accident, negligent care, age, or natural disease. Early deaths may be a manifestation of the higher test material intake by young animals on a mg/kg bw/day basis when compared to older animals (OECD, 2002a), so it is also important to consider the timing of all deaths.

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14.3.2  Clinical Observations General observations of all animals should be made daily. These clinical observations often entail recording any mor­ talities or obvious abnormalities observed while the animal remains in its cage. The test guidelines require that more expanded clinical observations be conducted weekly out­ side the home cage. These observations assess posture, gait, and reactivity to various stimuli, and thus provide a screen for neurotoxicity (Ross et al., 1998; Ross, 2000). They also include palpation of the animal for any masses (Keller and Banks, 2006). Clinical signs can be important indicators of target organ toxicity and not simply signs of general well-being (Keller and Banks, 2006). For example, cyanosis, flushing, and weakness may indicate toxicity to the cardiac system, while effects on the autonomic nervous system may be manifest as piloerection and lacrimation. It should be noted that vomiting is a natural occurrence in dogs (Haggerty et al., 2007). Consequently, potential effects on growth and changes in pharmacokinetics and test material intake due to vomiting of the test diet should be considered. The U.S. EPA test guideline also requires that a func­ tional observational battery (FOB) and motor activity testing be conducted toward the end of the first year of exposure (U.S. EPA, 1998a). In the FOB, the rat is system­ atically observed in its home cage, as it is being removed from its cage and held in the technician’s hand, and in an open field (Ross, 2000). The FOB assesses and quanti­ fies unconditioned behavior by evaluating neuromuscular function (gait, posture, grip strength), vestibular function (righting response, ataxia), sensorimotor responsiveness (response to crude stimuli), excitability state (reactivity), autonomic signs (pupil response, lacrimation, salivation, diarrhea), and tremors/convulsions (Moser, 2007).

14.3.3  Body Weight, Body Weight Gain, and Food Consumption An important endpoint in toxicology studies is body weight. Body weight gain is also measured to account for the use of young adult animals that are continuing to develop at the beginning of chronic toxicity studies (Keller and Banks, 2006). Food consumption is also recorded, as well as water consumption if the test material is adminis­ tered in the drinking water. Body weight and food or water consumption data are used to calculate the ingested dose. These endpoints are often affected at lower doses than other endpoints in chronic toxicity studies and thus are important for establishing the LOAEL in many cases. Changes in body weight, body weight gain, and food consumption are generally considered as indicators of gen­ eral systemic toxicity (Keller and Banks, 2006). Changes in body weight could signal a disturbance of metabolic,

Chapter | 14  The Assessment of the Chronic Toxicity and Carcinogenicity of Pesticides

hormonal and/or homeostatic mechanisms, or could be related to a wasting disease such as cancer or renal failure or even pain and distress. Alternatively, changes in body weight may be related to food intake; thus, any evaluation of body weight must be conducted in conjunction with an analysis of food consumption. This is of critical importance when a reduction in food consumption occurs as a result of decreased palatability of the test diet due to the presence of the test material, thus resulting in a lower body weight gain relative to control. In this situation, the effect on body weight should not be considered as a direct toxic effect of the pesticide. A lower body weight gain that is related to a palatability issue is often evidenced in the first 2 or 3 weeks of the study and may dissipate as animals develop a tolerance for the dietary admixture and start to grow at rates comparable to controls. However, animals may not recover from this initial delay in growth and differences in body weight with respect to controls may persist over the course of the study (OECD, 2002a). Thus, it is important that the variations in body weight/body weight gain and food consumption over the entire study period, not just the effects at termination, be taken into consideration. Other issues that demonstrate the importance of con­ sidering changes over the entire study period are that growth slows or stops in the latter stages of an animal’s lifespan and that older animals may exhibit weight loss due to age-related lesions (OECD, 2002a). In addition, animals that consume excessive diet and gain more weight tend to exhibit early onset of various degenerative dis­ eases and spontaneous tumors leading to shorter lifespans (Rhomberg et al., 2007), while restriction of food intake leads to increased survival (OECD, 2002a). Food conversion efficiency, calculated as the ratio of the amount of food consumed to the gain in body weight over a particular time interval, is another useful endpoint to evalu­ ate when assessing body weight and food consumption. Occasionally, treatment-related effects on food consump­ tion may not be readily apparent when comparing absolute food intake values, but treated animals, while consuming comparable amounts of diet as controls, are not growing at the same rate and are thus not as efficiently converting con­ sumed diet into useful nutrients and energy when compared to controls. Such an effect could be the result of impaired metabolic processes resulting from pesticide exposure or could be caused by a possible interaction of the test article with one or more essential nutritional elements in the diet. In situations where the food conversion efficiency is unaf­ fected in the presence of decreased body weights, reduc­ tions in food consumption are not likely the primary cause of the delayed growth (Keller and Banks, 2006). In addition to palatability in diet, important factors to con­ sider when assessing food consumption include food spillage and any condition that may cause eating to be uncomfortable, such as stomach ulcers. Reduced food consumption lead­ ing to weight loss may impair immune function and reduce

469

resistance to disease (OECD, 2002b). Young animals have a higher metabolic activity and growth rate than adult animals, which results in greater food consumption per unit of body weight. Increased food consumption on its own could be a manifestation of the animal’s attempt to compensate for a nutrient deficiency (Rhomberg et al., 2007). Water intake is not usually recorded in chronic toxi­city studies unless the test compound is administered in the drinking water. As with the dietary route, palatability of the drinking water may be an issue. Decreased water consump­ tion may result in reduced food intake (OECD, 2002b) and will have subsequent effects on urinalysis parameters.

14.3.4  Ophthalmoscopic Examination Indirect and slit lamp examinations should be conducted, and the administration of a mydriatic solution prior to examina­ tion can assist in viewing the deeper structures of the eye through the dilated pupil (Keller and Banks, 2006). For ocu­ lar toxicity, the pigmented eye of the dog is generally con­ sidered to be a better model of the human eye than the eye of the albino rat (Keller and Banks, 2006).

14.3.5  Clinical Pathology The clinical pathology portion of chronic toxicity studies consists of hematology, clinical chemistry, and urinalysis evaluations. Due to the comparable biology among mam­ mals, many of the examined parameters in laboratory ani­ mal testing are the same as those that are routinely tested for in humans (Hall and Everds, 2008). The battery of tests outlined in the guidelines can be considered as a screen that may signal the requirement for additional testing for more specific effects (Keller and Banks, 2006). Unless oth­ erwise specified, the following information on the interpre­ tation of clinical pathology results is obtained largely from comprehensive summaries provided by Smith et al. (2002), the OECD (2002b), and Hall and Everds (2008).

14.3.5.1  Hematology The hematological parameters that are normally assessed in chronic toxicity studies are outlined in Table 14.2. These parameters provide useful information on the balance between the production and destruction of cells of the cir­ culatory system as well as the clotting ability of the blood. Changes in these parameters could reflect an anemic con­ dition, leukemia, or inflammation. However, alterations in a single hematological parameter do not necessarily indi­ cate a toxic response unless accompanied by histological changes in the bone marrow or spleen (WHO, 1990). (a)  Red Blood Cell Parameters Reduced levels of circulating red blood cells and other red cell parameters (e.g., hemoglobin and hematocrit) usually

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indicate a form of anemia. These reductions may be caused by blood loss or by decreased production or increased destruction of red blood cells. Anemia may be classified as regenerative or nonregenerative. Regenerative anemia may be caused by hemorrhage or destruction of red blood cells and is accompanied by an increase in reticulocytes (immature red blood cells) and possibly higher mean cor­ puscular volume (MCV) and lower mean corpuscular hemo­ globin concentration (MCHC) values. Pesticides may cause increased destruction of red blood cells by directly damag­ ing red blood cell membranes or by oxidizing hemoglobin, which in turn leads to the formation of Heinz bodies with subsequent development of methemoglobinemia. Increased immune responses brought about by exposure to a chemi­ cal may also result in the destruction of red blood cells. Reticulocytosis is not observed in nonregenerative anemia, in which the bone marrow does not respond with a com­ pensatory increase in red cell production as it does in regen­ erative anemia (Bloom and Brandt, 2001). Nonregenerative anemias arise from reduced erythrocyte production in the bone marrow, which may be caused by chronic inflamma­ tion, impaired renal or hepatic function, hormonal imbal­ ances, or direct toxicity to bone marrow stromal cells. In these cases, the MCHC and MCV remain unaffected by exposure to the pesticide. Inadequate levels of folate or impaired DNA synthesis may result in a type of nonregener­ ative anemia (megaloblastic) in which mature red blood cells have undergone too few cell divisions and are larger than normal. Leukemia is another common cause of nonregenera­ tive anemia resulting from the displacement of stem cells in the bone marrow by cancerous cells. Dehydration may result in increased red blood cell counts, hemoglobin concentra­ tion, and hematocrit. (b)  White Blood Cells Handling and other activities such as blood sampling may frighten animals and cause elevations in white blood cells. Other causes include leukemia, lymphoma, infection, and necrosis. Decreases in white blood cell counts may be caused by overwhelming inflammation, peripheral leuko­ cyte destruction, bone marrow toxicity, loss of lymph, and stress (Lanning, 2006). In addition to total white blood cell counts, differen­ tial counts are performed to identify the type of leukocyte (i.e., neutrophils, lymphocytes, eosinophils, and basophils) that may be affected by treatment. It is recommended that focus be placed on absolute differential counts as relative cell counts are not considered to be useful. Identification of the type of white blood cell that is affected can help eluci­ date the cause of the change in total white blood cell count. For instance, increased neutrophils may occur in response to fear, excitement, or exercise or in association with hem­ orrhage, hemolysis, inflammation, or infection, while a decrease could occur following exposure to a cytotoxic

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chemical. Decreased blood lymphocyte counts could be due to a suppression of the immune system, significantly reduced food intake, stress, or exposure to corticosteroids. On the other hand, increases in lymphocyte counts may result from exposure to a compound that stimulates the immune system. Chronic inflammation can give rise to increases in lymphocytes and/or monocytes. Other causes of elevated monocytes include hemolytic anemia and tis­ sue death associated with tumor development or hypersen­ sitivity, while increases in eosinophils may be caused by hypersensitivity or parasitic infestation. Basophil counts are rarely affected by chemical exposure. (c)  Clotting Ability Clotting ability is assessed through the measurement of plate­ let levels and clotting times, the latter of which requires large blood volumes and thus may not be assessed in all studies (Keller and Banks, 2006). Increases in clotting times (pro­ thrombin time or activated partial thromboplastin time) may occur when absorption of vitamin K is compromised or when liver function (and consequently the production of most of the clotting factors) is severely affected. Thrombocytopenia usually occurs as a secondary effect to hemolytic anemia, hemorrhage, or infection, each of which stimulates the pro­ duction of cells, including platelets, in the bone marrow. Hemorrhage, if extensive, can also result in decreased plate­ lets. Decreases in platelets can translate to significant health consequences for the animal (Keller and Banks, 2006).

14.3.5.2  Clinical Chemistry A multitude of parameters are evaluated in the clinical chemistry portion of chronic toxicity studies conducted with pesticides (Table 14.2). The results of the clinical chemistry analysis provide valuable information related to metabolism of carbohydrates, proteins, and lipids, and the function of several organ systems, including the kidney, liver, muscle, and heart. (a)  Macronutrients (Lipids, Glucose, and Protein) Perturbations in cholesterol and triglycerides, commonly seen in chronic toxicity studies, usually reflect changes in food consumption and body weight or metabolic imbal­ ances resulting from liver dysfunction or hormone altera­ tions. Increases in triglyceride and cholesterol levels can be caused by hypothyroidism and diabetes mellitus, while increased cholesterol levels alone may be an indication of nephrotoxicity or cholestasis. Glucose concentrations may increase in nonfasted and moribund animals, in response to fear or pain, and with con­ ditions such as diabetes mellitus, pancreatitis, and hyper­ adrenocorticism. Lower glucose concentrations may result from breakdown by erythrocytes following the collection of

Chapter | 14  The Assessment of the Chronic Toxicity and Carcinogenicity of Pesticides

Table 14.2  Hematological, Clinical Chemistry, and Urinalysis Parameters Normally Evaluated in Chronic Bioassaysa Hematology

Red blood cell count Hemoglobin concentration Hematocrit Mean corpuscular volume Mean corpuscular hemoglobin Mean corpuscular hemoglobin concentration White blood cell count Differential leukocyte count Platelet count Clotting potential (e.g., prothrombin time or activated partial thromboplastin time)

Clinical chemistry

Potassium Sodium Calcium Phosphorus Chloride Glucose Total cholesterol Urea nitrogen Creatinine Total protein Total bilirubin Albumin More than two hepatic enzymes (such as alanine aminotransferase, aspartate aminotransferase, alkaline phosphatase, sorbitol dehydrogenase, or gamma glutamyl transpeptidase) Reticulocyte countb Bone marrow cytologyb Other measurements as appropriatec

Urinalysis

Appearance Volume Osmolality or specific gravity pH Protein Glucose Blood/blood cells

a

Adapted from OECD (1981a, 2009a) and U.S. EPA (1998a). Recommended to be evaluated if an effect on the hematopoietic system is evident. c Other measurements to be evaluated (e.g., triglycerides, hormones, methemoglobin, cholinesterases) if there is a known or suspected effect on that parameter. b

blood samples, decreased food consumption, malabsorption, and hepatic disease. Total protein, albumin, and globulin levels are mea­ sured. Dehydration results in increased protein levels, with albumin and globulins changing proportionately. Decreased protein concentrations may be observed following a period of decreased food consumption or may result from decreased synthesis (e.g., liver toxicity and poor digestion

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or absorption) or increased loss (e.g., hemorrhage and renal toxicity). Decreased albumin can be an indicator of acute inflammatory conditions, whereas increased globulins may result from increased synthesis following antigenic stimula­ tion leading to inflammation. Low albumin/globulin ratios are observed if albumin is selectively lost (e.g., glomerular disease) or not produced (e.g., hepatic disease). (b)  Indicators of Hepatotoxicity There are several parameters in the battery of clinical chem­ istry tests that are indicators of hepatic toxicity. These include a variety of enzymes (e.g., alanine aminotransferase, aspartate aminotransferase, sorbitol dehydrogenase, gluta­ mate dehydrogenase, and lactate dehydrogenase) that can be released from hepatocytes into circulation following membrane damage. Alanine aminotransferase (ALT) is con­ sidered to be the most specific of these enzymes to indicate hepatocellular injury, whereas increased circulating levels of aspartate aminotransferase (AST) or lactate dehydrogenase also may reflect muscle injury. Increased levels of ALT and AST may be caused by corticosteroids and anticonvulsants. Serum creatinine phosphatase activity is primarily a marker for toxicity to the skeletal muscle. Biliary toxicity is evaluated through the measurement of other enzymes, such as alkaline phosphatase (ALK) and gamma-glutamyl transferase (GGT), as well as bilirubin concentrations. As with the aminotransferases, ALK lev­ els can be increased following exposure to anticonvulsants and corticosteroids. Cholestasis will result in increased levels of circulating bilirubin, ALK, and GGT (although GGT is not a sensitive indicator of cholestasis in rats and mice). Decreasing levels of bilirubin may occur as a result of increased breakdown of bilirubin following exposure to mixed function oxidase inducers. Elevated levels of ALK may also indicate an effect on bone formation. Normally, reduced levels of hepatic enzymes are not considered adverse toxicological outcomes. However, decreases in serum ALT and AST activity may be observed when the test chemical has an effect on vitamin B6, which acts as a catalyst for reactions involving aminotransferases. (c)  Indicators of Renal Toxicity Toxic effects on the kidneys are revealed in the clinical chemistry analysis through the measurement of blood urea nitrogen (BUN) and creatinine, which often increase con­ comitantly in response to kidney injury. Elevated levels of BUN will also occur with dehydration. (d)  Minerals and Electrolytes A number of minerals and electrolytes are also assessed in the standard clinical chemistry evaluation. Often, small changes in these parameters can be an indication of toxi­ city. Although uncommon, increased calcium levels may occur in association with hyperparathyroidism or renal

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disease. Lower calcium levels may also be caused by renal disease, as well as hypoparathyroidism and pancreatitis. Much of the calcium circulating in the blood is bound to albumin, so changes in these two parameters are often synchronous. Elevations in serum inorganic phosphorus concentra­ tion may result from a decreased glomerular filtration rate. Reductions in phosphorus levels often arise following decreased food intake. Serum sodium and chloride levels will both rise in response to dehydration and fall follow­ ing fluid loss through the gastrointestinal tract or kidney or through exposure to diuretics. The diet plays an important role in the regulation of serum potassium concentrations, which rise in response to metabolic acidosis and fall in response to severely depressed food intake. An increase in potassium levels may also indicate renal disease.

14.3.5.3  Urinalysis The parameters analyzed in the urinalysis portion of chronic toxicity studies (Table 14.2) generally provide an indication of renal function but also yield information related to other organ systems. However, urinalysis is often considered “an imprecise tool” (Smith et al., 2002) which uses “crude [col­ lection] procedures” (Keller and Banks, 2006). Changes in urine volume and specific gravity, which frequently occur inversely to each other, often indicate an effect on the kidney’s ability to concentrate urine. Chronic renal disease can impair the kidney’s ability to concentrate urine, resulting in increased volume and decreased specific gravity and leading to excess urea in the blood. Although limited in value, urinary pH can be used to assess acid/base balance. The pH of the urine may be affected by diet (high protein diets tend to lower pH) or ammonia released by bacteria in samples collected overnight. Analysis of urine sediment may reveal the presence of renal tubular cells, an indicator of kidney disease. Excess protein in the urine may arise from injury to glomerular membranes or renal tubules, hemorrhage, inflammation, or as a consequence of progressive nephropathy, which is common in aging rats. Elevated urinary glucose concen­ trations may follow injury to the renal proximal tubules or may be due to diabetes mellitus, although this is not a common finding in toxicology studies. Trauma, inflam­ mation, or neoplasia of the kidney, ureters, or bladder are likely causes of the presence of blood in urine. Possible contamination of urine samples by hair and dust or even food should also be taken into consideration when evaluat­ ing urinalysis results (Keller and Banks, 2006).

14.3.6  Organ Weight Absolute and relative (to body weight and/or brain weight) organ weight data can be used in conjunction with clinical

Hayes’ Handbook of Pesticide Toxicology

pathology and other necropsy data to assess specific tar­ get organ toxicity. However, organ weights can easily be affected by the method and timing of sacrifice as well as altered growth. Fasting prior to sacrifice can decrease the variability of organ-to-body weight ratios. The rela­ tive weight of the brain, in particular, is highly dependent on changes in body weight since absolute brain weight is relatively insensitive to body weight effects (Keller and Banks, 2006; OECD, 2002a). For this reason, organ-tobrain weight ratios become important indicators of organ weight effects when body weight is affected (WHO, 1990). However, treatment-related effects on organ weight should not be overlooked when effects on body weight are evident (Keller and Banks, 2006). Adaptation to stress or metabolic overload may lead to increases in relative organ weights, which often return to normal once these stressors are removed through the cessation of dosing (WHO, 1978). Cyclical reproductive processes can cause changes in uterine or ovarian weight (Keller and Banks, 2006).

14.3.7  Macroscopic and Microscopic Pathology Macroscopic (gross) and microscopic (histo-) pathology, examined in a wide range of organs and tissues in chronic toxicity studies, are primary indicators of specific organ toxicity. The pathological examination identifies morpho­ logical changes in tissues at the gross and microscopic levels, describes the type of lesion (e.g., inflammatory, degenerative, or disturbances of growth), and determines the cause of morbidity and death (Frame and Mann, 2008). Standardized classification and terminology should be used in the pathological examination as much as possible to ensure comprehensibility and allow for comparison to findings in other studies (Keller and Banks, 2006). Histopathology is considered to be a more sensi­ tive marker of organ toxicity than organ weights (OECD, 2002a,b) and often forms the basis for establishing the LOAEL (WHO, 1990). Effort should be made to corre­ late any gross findings with microscopic lesions. In addi­ tion, an attempt to delineate the cause of any findings noted upon necropsy should be made, including those that occur with low incidence (WHO, 1978). Lesions may be due primarily to exposure to the pesticide, may be second­ ary to another toxic effect of the pesticide, may simply be spontaneous occurrences, or could be due to postmortem changes. For example, animals that die prior to sacrifice often will be reported as having congestion in numerous organs upon necropsy due to blood settling in these organs as well as an accumulation of gas in their intestines due to autolysis (Keller and Banks, 2006). Correlating necropsy findings with in-life observations is an integral step in the interpretation of the results of the pathology examination.

Chapter | 14  The Assessment of the Chronic Toxicity and Carcinogenicity of Pesticides

It is important to differentiate between spontaneous and/or age-associated lesions and those lesions induced by treatment. Exposure to a pesticide may affect the devel­ opment of age-related lesions by (1) increasing the inci­ dence, (2) increasing the severity, and/or (3) accelerating the development. While usually an increase in the inci­ dence of a particular lesion is of concern, exposure to a test chemical may appear to decrease the incidence of cer­ tain lesions commonly observed in aged animals. This may occur when, for instance, survival is reduced such that agerelated lesions have not had the opportunity to develop or decrements in body weight gain provide a protective effect from age-related conditions (OECD, 2002b).

14.3.8  Additional Endpoints Although not required in current chronic toxicity test­ ing guidelines, additional emphasis is being placed on the testing of neurological (beyond the FOB and motor activ­ ity testing), immunological, and reproductive endpoints in repeated-dose studies as screening tools to assess if further, more elaborate testing in these areas is required (OECD, 2002a). A need has been identified to assess the potential for pesticides to target the nervous system using studies of longer duration beyond the 3 months normally used for neurotoxicity testing (Moser, 2007). Cholinesterase activity in the blood, erythrocytes, and/ or brain often is examined with certain pesticides known to affect this enzyme, such as the organophosphates and car­ bamates. Methemoglobin, formed when the heme iron of hemoglobin is oxidized (Lanning, 2006), is an additional parameter that may be measured in the hematology workup. Microsomal enzyme induction is often tested to deter­ mine if pesticide exposure can induce hepatic cytochrome P450 enzymes, as well as associated mixed function oxi­ dase activity and other enzymes. If such inductions lead to increased metabolism of the pesticide, alterations from control animals noted early in chronic toxicity studies may dissipate over the course of the study as the animals begin to metabolize the test material more efficiently (Smith et al., 2002). Circulating levels of hormones, such as thyroid hormones, also may be measured to assist in characterizing the mode of action or to identify a point of departure for a critical toxicological effect mediated by hormonal changes. Cardiovascular endpoints also may be assessed through the evaluation of electrocardiography and blood pressure, although such assessments are usually conducted in nonro­ dents due to their larger size (Keller and Banks, 2006). The endpoints indicated previously are by no means an exhaustive list of parameters that may be assessed in chronic toxicity studies. The nature of the pesticide toxi­ city and the characteristics of the exposed population should dictate which endpoints beyond those described in the standardized test guidelines are examined.

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On a final note, many chronic toxicity studies employ a recovery experiment, in which the reversibility of toxic effects following cessation of exposure is assessed. Such information is essential to understanding regeneration and repair mechanisms, as well as the potential for cumulative toxicity or progression of toxic effects (Barile, 2008).

14.3.9  Overall Assessment Effects observed in chronic toxicity studies should be considered with respect to both their statistical and their biological significance. One should not rely on statistical significance alone, especially when only a few animals are assigned per group as is the norm in nonrodent stud­ ies. When only a few individuals in a small group respond to treatment, a wider dispersion of the data may occur that could reduce the sensitivity of statistical tests. The relation to treatment, adversity, and human relevance are all consid­ ered when determining the point of departure for chronic endpoints. Most parameters are assessed repeatedly over the course of a chronic toxicity study. These repeated analyses pro­ vide useful information on the time course of development of toxicological changes, which could help elucidate the cause for the observation or the process involved/affected. For example, a hematological change that occurs early in a study is more likely to be due to hemorrhage or hemoly­ sis rather than a direct effect on red blood cell production or an immune-mediated response, given the relatively long lifespan of the red blood cell (Smith et al., 2002). Repeated analysis may also reveal adaptation to the effects of expo­ sure to the chemical. Finally, toxicity may be enhanced over time if repeated exposure leads to induction of meta­ bolic processes, which then results in elevated levels of a toxic metabolite, or it may be reduced over time if the par­ ent is the primary toxicant (Keller and Banks, 2006). Differences between the sexes are evaluated, with the caveat that male rats tend to metabolize xenobiotics faster than female rats (OECD, 2002b). Results from metabolism studies are useful in identifying potential target organs as well as differences in toxicokinetics and toxicodynam­ ics between the sexes. The relationship between response and duration of exposure is also assessed by compar­ ing the results from longer term studies with those from subchronic studies for a given species and/or through the examination of interim sacrifice groups and animals that die or are sacrificed in moribund condition before study termination (WHO, 1990). The relevance of effects noted in chronic animal bioas­ says to human exposures is also taken into consideration. For example, not all sex differences are necessarily rele­ vant to humans given the disparity in metabolic capa­city between male and female rats noted previously (OECD, 2002b). However, information pertaining to metabolism or

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toxic effects of a pesticide in humans is rarely available. Unless a species-specific effect is clearly demonstrated with supportive data, most toxic responses observed in the animal models are assumed to be relevant to humans (OECD, 2002a). When dissimilar effects occur in different species, the most sensitive species is used for risk assess­ ment purposes unless data are available to indicate that another species is more relevant to the human health risk assessment (WHO, 1990). For this purpose, the Human Relevance Framework (Meek et al., 2003) developed for carcinogenic endpoints (described in Section 14.4) can be applied to noncancer, chronic toxicity endpoints as well (Cohen et al., 2004; Seed et al., 2005). The effects observed in the chronic toxicity studies conducted with pesticides are not analyzed in isolation. First, findings within the chronic study are correlated with other endpoints within the same study to assist in the inter­ pretation and explanation of toxic responses. Information from the chronic toxicity study is then integrated with results from subchronic and other toxicity studies, as well as toxico- and pharmacokinetic data. To assess the level of concern for the toxic responses observed, a weight-ofevidence approach is taken, which relies heavily on scientific judgment and the assemblage of data from all available toxicity studies for a particular pesticide.

14.4  Assessment of carcinogenic potential The identification of a pesticide as a potential carcinogen involves the use of internationally recognized test guide­ lines, protocols, and approaches that consider, among other things, noncarcinogenic endpoints and statistical methods, as well as the mode of carcinogenic action in animals and its relevance to humans. Once the carcinogenic hazard of a pesticide is characterized, the appropriate risk assessment method must be defined and applied, as outlined further in Section 14.5.

14.4.1  Overall Approach The carcinogenic potential of a pesticide is determined through an evidence-based approach using the results from in vivo carcinogenicity bioassays in at least two species conducted according to the guidelines described previously as well as the results from in vitro and in vivo genotoxi­city studies and other sources of information, such as epide­ miological studies, structure–activity relationships, and mechanistic or mode of action studies (Farland et al., 2006; Health Canada, 1991). Of primary importance is the histopathological exami­ nation. In rodent carcinogenicity bioassays, this exami­ nation identifies hyperplastic, dysplastic, and neoplastic

Hayes’ Handbook of Pesticide Toxicology

lesions and classifies neoplasms as benign or malignant and as primary or metastatic (Frame and Mann, 2008). The data are assessed to determine the presence of tumors, the number of different tumor types, and the number of animals affected (Barile, 2008). The data are scrutinized to identify the potential for the induction of rare tumors as well as the earlier or increased induction of spontaneous, commonly observed tumors. In addition to the presence or absence of neoplasia, effects on survival and body weight gain should also be considered, along with indications of a preneoplas­ tic response, time to tumor onset, and the presence of mul­ tiple tumors at various sites (Williams et al., 2008). Concern regarding potential carcinogenesis in humans is heightened if (1) tumors are observed in more than one species; (2) a rare tumor is observed; (3) tumor develop­ ment is noted in both sexes; (4) tumors develop at several sites or by more than one route of exposure; (5) the tumor response in one species is repeated in more than one study; (6) progression from preneoplastic to benign to malignant occurs; or (7) a positive response in genotoxicity studies is also observed (OECD, 2002b; U.S. EPA, 2005a; Williams et al., 2008). While concern for benign tumor develop­ ment is tempered if the type of lesion does not progress to a malignant tumor (U.S. EPA, 2005a; Williams et al., 2008), benign and malignant tumors are generally regarded equally with respect to concern for human health, and risks to both from pesticide exposure are assessed (OECD, 2002b), unless there is evidence to indicate that the benign tumor would have no significant impairment to human health (U.S. EPA, 2005a). The degree of concern may be lessened if tumor devel­ opment is observed only following exposure via a route having little relevance to humans, if the species demon­ strating a positive response is a poor model for human assessment due to a difference in metabolic processes, or if a positive response is observed at doses that were not well tolerated by the animals (i.e., at doses that exceeded the MTD) (OECD, 2002b). In addition to dose level selection (described in Section 14.2.3), one of the most critical aspects in defining the adequacy of a carcinogenicity study is survival at study termination. This is of particular importance when results from the study indicate that a pesticide is negative for car­ cinogenicity. A carcinogenic effect may be missed when a significant number of early deaths occur, since it may be that an insufficient number of animals lived long enough to develop tumors, thus reducing statistical power (Rhomberg et al., 2007). For a negative test to be considered accept­ able, the OECD (1981b) recommends that survival of all groups must be at least 50% at the end of the study, while the U.S. EPA (1998b) stipulates that survival should not fall below 50% at 15 months for studies conducted with mice and at 18 months for studies using rats, or below 25% at study termination.

Chapter | 14  The Assessment of the Chronic Toxicity and Carcinogenicity of Pesticides

The current classification system for carcinogens in the United States (outlined in U.S. EPA, 2005a) uses a set of descriptors to classify pesticides based on their likeli­ hood to cause cancer in humans. These descriptors include “carcinogenic to humans,” “likely to be carcinogenic to humans,” “suggestive evidence of carcinogenic potential,” “inadequate information to assess carcinogenic poten­ tial,” and “not likely to be carcinogenic to humans.” The current and previous classification systems used in the United States are outlined in Table 14.3. The assignment of a particular descriptor is based on the strength of the animal evidence and epidemiological data. When the carci­ nogenic response differs for different dose levels or expo­ sure routes, more than one descriptor may be assigned to a pesticide. Generally, potential risks to humans at expo­ sure levels of interest are estimated using dose–response assessments for pesticides considered “carcinogenic to humans” and “likely to be carcinogenic to humans” (U.S. EPA, 2005a). When a pesticide is classified as “sugges­ tive evidence of carcinogenic potential,” a dose–response assessment is not normally conducted, but the data may be used to establish a need for further testing or to estimate uncertainty in the risk characterization (U.S. EPA, 2005a). Health Canada’s Pest Management Regulatory Agency (PMRA) does not use a classification system, but instead evaluates pesticides on a case-by-case basis to determine carcinogenic potential.

14.4.2  Statistical Considerations The biological significance of a carcinogenic effect should be considered along with statistical significance when determining the carcinogenic hazard of a pesticide. Several factors need to be considered in selecting appropriate sta­ tistical tests to determine statistical significance of a tumor response. These factors include, but are not limited to, the normal background incidence of the tumor, intercur­ rent mortality, survival rates in the different dose groups, and whether tumors can be considered incidental or fatal (Haseman, 1984; Jacobs, 2006). Benign and malignant neoplastic lesions in tissues of the same histological ori­ gin may be combined for statistical assessment of tumor response under certain circumstances, such as when hyper­ plasia cannot be easily differentiated from benign neopla­ sia in a histological examination (McConnell et al., 1986). Tumor response in male and female animals should not be combined for statistical assessment (Jacobs, 2006). The unit of analysis is normally based on the number of ani­ mals exhibiting a tumor response as opposed to the number of tumors (Jacobs, 2006). Statistical tests used to assess the probability that tumor development is not due to chance include trend tests, pairwise tests, and, if required, time-adjusted or survivaladjusted tests (Jacobs, 2006; Velasquez et al., 1995). Trend

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tests, such as the Cochrane–Armitage trend test, examine whether incidences in all dose groups considered together increase as a function of dose, while pairwise tests, such as the Fisher exact test, assess the tumor response in each dose group individually against the tumor response in the control group (Jacobs, 2006; Velasquez et al., 1995). These tests do not take into account intercurrent mortality (Haseman, 1984). Adjusting for intercurrent mortality, or deaths other than those resulting from the tumor of interest, is essen­ tial to reducing bias caused by differences in survival rate among dose groups (Lin, 2003). Time-adjusted or survivaladjusted tests, such as life table analyses, are used when there are differences in survival rates among the dose groups (Jacobs, 2006). Information on whether death was caused by a tumor is not always available for rodent car­ cinogenicity studies. The use of interim sacrifice groups may help elucidate time-to-tumor onset. Poly-k type tests, which are modifications of the Cochrane–Armitage trend test, consider dose–group differences in intercurrent mor­ tality and are used when cause of death information is not available (Lin, 2003). Life table analyses assume that all tumors appearing in animals that die before study ter­ mination are fatal (Haseman, 1984). Peto type tests (Peto et al., 1980) may be used if information on tumor lethality is available. The death-rate method, the onset-rate method, and the prevalence method are used for fatal tumors, mortality-independent tumors (observable without nec­ ropsy, such as skin and mammary gland tumors), and inci­ dental tumors, respectively (Lin, 2003).

14.4.3  Mode of Action During the assessment of a pesticide for carcinogenic­ ity, the mode of action for potential carcinogens is taken into consideration. The term “mode of action” refers to a series of key events describing interaction of the pesticide with the organism at the cellular level and the subsequent structural and functional changes that result, and is different from “mechanism of action,” which is a more detailed description of the process of interaction down to the molecular level (U.S. EPA, 2005a). Key events are defined as “measurable events that are critical to the induc­ tion of tumors as hypothesized in the postulated mode of action” (Sonich-Mullin et al., 2001). In general, pesticides may cause cancer through a genotoxic mode of action, in which the formation of cancer is due to the pesticide react­ ing directly with DNA of the organism, or through a non­ genotoxic mode of action, in which cellular proliferation in tissues arises through a means that does not involve inter­ action with genetic material. Whether in different tissues or the same tissue, a pesticide may act through more than one mode of action (U.S. EPA, 2005a). Mode of action data are included in the overall weight of evidence analysis

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Table 14.3  U.S. EPA Carcinogenicity Classification for Pesticidesa Year

Classification/descriptor

Definition

1986

Group A—Human carcinogen

Sufficient evidence from epidemiologic studies

Group B—Probable human carcinogen

Sufficient evidence from animal studies and: limited evidence from epidemiologic studies (Group B1), or l �������������������������������������������������� inadequate evidence or no data from epidemiologic studies (Group B2) l

1996

1999

2005b

a

Group C—Possible human carcinogen

Limited evidence in animals in the absence of human data

Group D—not classifiable as to human carcinogenicity

Inadequate human and animal evidence or no data available

Group E—Evidence of noncarcinogenicity for humans

No evidence in

Known/likely

Data are adequate to convincingly demonstrate carcinogenic potential for humans

Cannot be determined

Data are suggestive, conflicting, or limited in quantity and are thus inadequate to convincingly demonstrate carcinogenic potential for humans

Not likely

Evidence is satisfactory for deciding that there is no basis for human hazard concern

Carcinogenic to humans

Convincing epidemiologic evidence demonstrating causality between human exposure and cancer

Likely to be carcinogenic to humans

Data are adequate to demonstrate carcinogenic potential to humans

Suggestive evidence of carcinogenicity, but not sufficient to assess human carcinogenic potential

Evidence from human or animal data is suggestive of carcinogenicity but is not sufficient for a conclusion as to human carcinogenic potential

Data are inadequate for an assessment of human carcinogenic potential

Data are judged to be inadequate to perform an assessment

Not likely to be carcinogenic to humans

Data are robust for deciding that there is no basis for human hazard concern

Carcinogenic to humans

Strong evidence of human carcinogenicity

Likely to be carcinogenic to humans

The weight of the evidence is adequate to demonstrate carcinogenic potential to humans but does not reach the weight of evidence for the descriptor “Carcinogenic to humans”

Suggestive evidence of carcinogenic potential

Evidence is suggestive of carcinogenicity (concern for potential carcinogenic effect in humans but data are judged not sufficient for a stronger conclusion)

Inadequate information to assess carcinogenic potential

Data are judged inadequate for applying one of the other descriptors

Not likely to be carcinogenic to humans

Data are considered robust for deciding that there is no basis for human hazard concern

����������������������������������������������������� at least two adequate animal tests in different species, or l �������������������������������������������������� both adequate epidemiologic and animal studies l

Adapted from U.S. EPA (2006). Multiple descriptors: more than one descriptor can be used when effect differs by dose or exposure route.

b

Chapter | 14  The Assessment of the Chronic Toxicity and Carcinogenicity of Pesticides

for carcinogenicity and are used to determine the type of cancer risk assessment (i.e., threshold or nonthreshold) to be conducted, if required. The risk posed by genotoxic car­ cinogens is usually assessed through a nonthreshold (lin­ ear) approach, whereas carcinogens demonstrated to act through a nongenotoxic mode of action generally undergo a threshold (nonlinear) risk assessment (further described in Section 14.5). By default, it is generally assumed in pesticide regula­ tion that a pesticide (or its metabolite) acts through a geno­ toxic mode of action in the formation of tumors. Evidence of positive interaction with DNA from genotoxicity studies, usually in vitro gene mutation and structural chromosome aberration assays and in vivo assays (U.S. EPA, 2005a), is a strong indication that the pesticide acts through a geno­ toxic mode of action. Structural similarity to chemicals known to operate through a genotoxic mode of action pro­ vides supporting evidence that the pesticide may behave in a similar manner (U.S. EPA, 2005a). In general, nongenotoxic carcinogens increase tumor development by enhancing cell division and/or inhibit­ ing apoptosis (Gregus and Klassen, 2001). Nongenotoxic modes of action must be fully elucidated and supported with acceptable data prior to acceptance by pesticide regu­ lators. The U.S. EPA Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a) describe a detailed frame­ work for analyzing the scientific information needed to assess a chemical’s carcinogenic mode of action, which is based on a framework developed by the International Programme on Chemical Safety (Sonich-Mullin et al., 2001). The framework includes the following steps to be applied to each tumor type under evaluation, although it is emphasized that the framework is not a “checklist of cri­ teria, but rather an analytical approach” (Sonich-Mullin et al., 2001): (1) description of hypothesized mode of action (including identification of key events); (2) discus­ sion of the experimental support for the hypothesized mode of action (including an assessment of causality taking into consideration the strength, consistency and specificity of the association, the dose–response concordance, tempo­ ral relationships, and biological plausibility and coher­ ence); (3) consideration of the possibility of other modes of action; and finally (4) conclusions about the hypoth­ esized mode of action (including the level of support for the hypothesized mode of action provided by the animal data, the relevance of the hypothesized mode of action to humans, and consideration of susceptibility of subpopula­ tions or life stages) (Sonich-Mullin et al., 2001; U.S. EPA, 2005a). In essence, the framework outlines an approach to evaluating how exposure to a pesticide may cause tumor development. When a hypothesized mode of action is not fully supported by the data, the framework allows for the identification of data gaps in the mode of action argument (Farland et al., 2006).

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14.4.4  Human Relevance In pesticide evaluation, the process through which tumors develop in animals is assumed to be relevant to humans unless data are provided to demonstrate otherwise or there is sufficient evidence for a particular tumor type that indi­ cates human irrelevance. For example, renal tumors aris­ ing from the binding of a pesticide to (alpha)2-globulin in male rats are not expected to occur in humans as humans do not produce this protein or an equivalent (Cohen et al., 2004). The relevance of the hypothesized mode of action to humans is included as part of the U.S. EPA’s animal mode of action framework under the final step in which conclusions about the mode of action are made (U.S. EPA, 2005a). An assessment of the sequence of key events iden­ tified in animals that are sufficiently supported by data is conducted. This assessment determines which key events may or may not be likely to occur in humans given their physiology compared to that of the test animals. An expanded framework (Meek et al., 2003) has also been developed to assess the human relevance of carcino­ genic modes of action. In addition to the analysis of the evidence provided to establish a mode of action in animals and the qualitative consideration of the applicability of the key events to humans, this expanded framework takes into consideration quantitative aspects (kinetic and dynamic factors) to determine if the animal mode of action is plau­ sible in humans (Cohen et al., 2003). Cohen et al. (2004) describe the application of this human relevance frame­ work through the presentation of specific examples. When use of the human relevance framework determines that the mode of action for tumor development in animals is not relevant to humans, a human health risk assessment is not required for that particular hazard.

14.5  Application to risk assessment and regulatory decision making To assess the probability of harm resulting from exposure to a pesticide, a risk assessment is conducted that takes into account the toxic effects that the pesticide may produce, the dose levels at which these effects may occur, and the magnitude of the potential exposure (OECD, 2002a). The approach to pesticide risk assessment varies depending on the mode of action through which a pesticide exerts a par­ ticular toxic effect. As indicated earlier, it is generally con­ sidered that there is a dose or concentration below which adverse effects will not occur (i.e., a threshold exists) for the majority of noncancer toxicity endpoints. For most types of cancer, the assumption is that there is some prob­ ability of harm at any level of exposure (i.e., no threshold exists). When sufficient evidence adequately supports a nongenotoxic mode of action, a threshold approach may be

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taken in the risk assessment for the carcinogenic hazard. Both threshold and nonthreshold approaches may be used when there are multiple modes of action for a particular pesticide (U.S. EPA, 2005a).

14.5.1  Chronic Toxicity Risk Assessment –   Noncancer Endpoints For chronic, noncancer effects, the point of departure, or “threshold,” below which toxicity is not expected to occur is determined. The NOAEL is defined as the high­ est dose that does not result in observable adverse effects in test animals, whereas the LOAEL is the lowest dose that causes an observable adverse effect. Some regula­ tory authorities use the term NOEL or no-observed-effect level, which can be defined as the highest dose at which there is an absence of any observable effect on measured endpoints in a study. Similarly, the LOEL, or the low­ est-observed-effect level, is the lowest dose that causes any change, whether adverse or not, distinguishable from control animals. The effect levels established in chronic toxicity studies can also be used to extrapolate to a bench­ mark dose. The benchmark dose method aims to define a point of departure that is more independent of study design than the determination of the NOAEL and LOAEL (U.S. EPA, 2008). WHO (1999) defines the benchmark dose as “the effective dose (or its lower confidence limit) that produces a certain increased incidence (or response) above control level (e.g., 1 or 5% of the maximum toxic response).” Modeling of the data is conducted to arrive at the benchmark dose. The benchmark dose method requires that the effects of interest are observed at several dose lev­ els but offers many advantages over the traditional estab­ lishment of a NOAEL as it takes into account the slope of the dose–response curve, the size of the study groups, and variability in the data (OECD, 2002b; WHO, 1999). The pertinent findings in the pesticide’s entire toxi­ cological database are considered when choosing the rel­ evant endpoint to use in chronic risk assessment. The lowest NOAEL in the most sensitive species is usually, but not always, selected. Several factors are considered when determining the appropriate effect level for use in chronic risk assessment, which includes the relevance of the toxic effect to humans and the adequacy/validity of the study from which the lowest effect level is derived. Differences in toxicokinetics between the most sensitive species and humans may render results in the sensitive species irrel­ evant to the prediction of human toxicity. Once selected, the NOAEL (or benchmark dose) from the most suitable study is divided by the appropriate assessment factor to estimate the maximum amount of pes­ ticide to which a human may be exposed daily over his/her lifetime without appreciable health risk. The assessment factor (also described by some regulatory authorities as an

Hayes’ Handbook of Pesticide Toxicology

uncertainty factor and/or a safety factor or a combination thereof) is a numerical adjustment applied to the point of departure to “arrive at a criterion or standard that is con­ sidered safe or without appreciable risk” (OECD, 2003). A detailed description of the assessment factor will not be provided here as approaches differ among regulatory bod­ ies, and consideration of the entire toxicology database is required when determining the appropriate magnitude of these factors. Briefly, the assessment factor addresses uncertainty inherent in the extrapolation from an animal species to humans and from variability in response among individuals within the human population, usually with the use of a 100-fold factor. The value of the assessment fac­ tor may be increased after taking into account quality of the data supporting the point of departure used for risk assessment, the quality of the entire toxicological dataset, the shape of the dose–response curve, and how well the study duration represents the expected duration of human exposure, among other things. It should be noted that the value of the assessment factor is also taken into con­ sideration when selecting the most appropriate endpoint for use in chronic risk assessment. For example, a higher NOAEL that warrants a larger assessment factor may result in greater protection than a lower NOAEL for which a smaller assessment factor is justified. Principles behind the selection of assessment factors are summarized in WHO (1999), U.S. EPA (2002), and Health Canada (2008). The estimated amount of pesticide to which a human may be exposed daily over his/her lifetime without appre­ ciable health risk may be called the acceptable daily intake (ADI) or the chronic reference dose when establishing safe levels in the diet. Generally, for nondietary scenarios such as workers exposed dermally and through inhalation, a margin of exposure approach is taken. In this approach, the ratio of the NOAEL for the critical toxicological effect to the theoretical, predicted, or estimated exposure level is calculated and then compared to the assessment factor. Exposure is not expected to result in health risks when this ratio exceeds the assessment factor.

14.5.2  Cancer Risk Assessment – Threshold Carcinogens When data corroborate the supposition that tumors arise through a nongenotoxic mode of action, a nonlinear or threshold approach to assessing the cancer risk is con­ sidered appropriate. It is presumed to be unlikely that nongenotoxic carcinogens pose a cancer risk to humans at exposure levels that do not cause any “key events” in the mode of action to occur. In these cases, a thresh­ old approach similar to the one described previously for chronic toxicity endpoints is applied; that is, a point of departure (NOAEL) for the carcinogenic mode of action is determined along with the suitable assessment factor to

Chapter | 14  The Assessment of the Chronic Toxicity and Carcinogenicity of Pesticides

ensure that an adequate margin exists between the lowest dose that results in the initiation of the key events in the carcinogenic mode of action and the anticipated lifetime human exposure.

14.5.3  Cancer Risk Assessment – Nonthreshold Carcinogens When data indicate that a pesticide acts through a geno­ toxic mode of action, or when there are inadequate data to indicate otherwise, a nonthreshold approach is applied to the cancer risk assessment. This is normally accomplished through linear low dose extrapolation which uses a mathe­ matical model to extrapolate the dose–response curve from the relatively high doses administered to animals into the low dose region, which is presumably more representative of human exposure levels. While several different mathe­ matical models are available for this purpose, currently the linearized multistage model is most widely applied (WHO, 1999). The purpose of low dose extrapolation is to provide a quantitative estimate of the cancer risk. The slope of the dose–response curve in the low dose region, generated through the linearized multistage model, represents the change in risk per increase of unit dose in the test spe­ cies. This slope is denoted as q1. The unit risk, defined as the increase in excess risk per unit dose, is the upper 95% confidence limit on this slope and is denoted by q1*. Dose scaling factors are then applied to the calculated unit risks to account for differences in metabolic rate between humans and the test species. These converted unit risks are multiplied by the anticipated pesticide exposure amortized over a lifetime to obtain the lifetime cancer risk values, which are estimates of the likelihood or probability of can­ cer. Risk management implications play into the regulatory decision regarding the acceptability of cancer risk. A life­ time cancer risk that is below one in a million, indicating that an individual exposed to the pesticide will have less than a one in a million chance of developing the cancer in question, generally is regarded as being negligible. Oftentimes, rodent carcinogenicity studies for pesti­ cides are conducted by the dietary route, while long-term exposure by the dermal and/or inhalation routes may be expected for workers handling the pesticide. In these cases, route-specific cancer risks cannot be calculated without a route-specific animal study. Therefore, the unit risks obtained from the dietary carcinogenicity studies are extrapolated to the cancer risk assessments for the dermal and/or inhalation routes of exposure. Scientific judgment must be exercised to ascertain the likelihood that the can­ cer observed following oral dosing would occur following other routes of exposure, taking into consideration poten­ tial for absorption, portal of entry effects, and probability of the pesticide reaching the target organ.

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14.6  Challenges While the practices governing the assessment of the chronic toxicity and carcinogenicity of pesticides are well established, recognized, and accepted worldwide, some limitations in the current testing paradigm do exist. Attempts are being made in the scientific community to address these limitations. In addition to the complexity surrounding dose selection in carcinogenicity discussed previously, exposure during various life stages and the assessment of the cumulative toxicity of several pesticides are two of the current issues facing pesticide regulators and international scientists in the field of toxicology.

14.6.1  Life Stages There is evidence that early life (e.g., prenatal and neona­ tal) exposure to carcinogens increases the susceptibility to cancer (which may manifest in childhood or later in life), particularly for those carcinogens that act via a mutagenic mode of action (Barton et al., 2005; U.S. EPA, 2005b). Early life exposure to a carcinogen does not necessarily lead to the development of tumors different from those observed following exposure later in life but results in an increased incidence of tumors or a reduced latency period (Barton et al., 2005). The OECD acknowledges that organ­ isms in early life stages may be more susceptible to car­ cinogenesis than the adult but cautions that there is only limited evidence that the carcinogenic potential of a chemi­ cal would not be detected in studies in which exposure began in early adulthood versus during the prenatal period (OECD, 1981c). However, a full assessment of children’s risk is limited currently by the lack of testing protocols that focus on hazards presented to early life stages (U.S. EPA, 2005a). While the current guidelines for chronic toxicity and carcinogenicity testing require that exposure to ani­ mals begins in early adulthood, studies using prenatal or neonatal animals may be recommended under special con­ ditions (U.S. EPA, 1998c). Assessment of children’s risk is further limited by the fact that the majority of the can­ cer epidemiology data considered robust enough to draw conclusions from arises from occupational settings which employ adults (Farland et al., 2006). Despite these limitations, risks to children are not ignored in the carcinogenic risk assessment of pesticides. Age-specific factors such as eating habits and body weight are considered in the derivation of lifetime exposure esti­ mates (Farland et al., 2006). The U.S. EPA (2005b) has developed guidance on potency adjustment for carcino­ gens known to act through a genotoxic mode of action. To account for the potential for early life exposures to make a greater contribution to the development of cancer later in life, adjustment factors of 10 and 3 are recommended to be applied to the cancer potency estimates (slope factors)

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derived from adult animals for exposure before 2 years of age and for exposures between 2 and 16 years of age, respectively (U.S. EPA, 2005b). On the opposite end of the spectrum, the elderly also may be more susceptible to the effects of pesticide expo­ sure due to their altered physiology and biochemistry (WHO, 1993). Dosing in rodent studies is continued into later life stages, but extending the dosing period may con­ found interpretation of some findings. As rodents age, they experience declining organ function, display reduced metabolic capa­city, and develop age-related diseases and tumors; consequently, the variability of the response to pesticides increases within the population (OECD, 2002b; WHO, 1987).

14.6.2  Cumulative Toxicity Exposure to more than one chemical may result in various types of interactions. These include potentiation (the toxicity of one substance is enhanced in the presence of another substance), additive interaction (the toxicity of two or more substances is equal to the sum of the individual effects), synergism (the toxicity of two substances is greater than the sum of the individual effects), and antagonism (the toxicity of two or more substances oppose one another) (Barile, 2006). Neoplasm promotion (enhancement of neoplastic development by a second agent given after an initiating carcinogen) or photochemical carcinogenesis (combined skin carcinogenicity of a chemical and ultravio­ let light radiation) may also occur (Williams et al., 2008). Currently, chronic toxicity and carcinogenicity are assessed for individual pesticides; the testing of mixtures is generally not conducted. The assessment of cumulative toxicity becomes complicated as there are different com­ binations of pesticide residues to which people may be exposed through the diet and through the workplace and the household uses of pest control products, not to mention exposures to other types of chemicals. The assessment of mixtures is also wrought with uncertainty, and most assess­ ments of mixtures are conducted with simple mixtures of two commercial chemicals (Farland et al., 2006). This is a challenge that all pesticide regulatory bodies are facing. However, it is the opinion of many in the international scientific community that the likelihood is low that toxic interactions from exposures to multiple pesticides with dis­ similar modes of action would occur at levels below the thresholds established for individual pesticides (Boobis et al., 2008; EFSA, 2008; VKM, 2008). For those pes­ ticides that exhibit a common toxic mode of action, the potential for adverse effects may be additive. Initiatives are in place to conduct cumulative risk assessments for those pesticides identified as having a common mechanism of toxicity, such as the organophosphates, N-methyl carba­ mates, triazines, and chloroacetanilides.

Conclusion It is the role of pesticide regulators, above all else, to pro­ tect human health. Modern risk assessment practices allow regulators to include long-term health effects and cancer in their evaluation of pesticides. As the science of human health risk assessment evolves, it is imperative that the tools utilized by pesticide regulators be refined continually to ensure that human health risk assessments for pesticides are maintained to the highest of standards. Authorities responsible for the regulation of pesticides, such as the U.S. EPA Office of Pesticides Program and Health Canada’s PMRA, remain intimately involved in, and often lead, the development of new risk assessment approaches in the international scientific community. The issues of life stages and cumulative toxicity continue to be focal points of international deliberations dealing with pesticide regulation. The current scrutiny with which pesticides are assessed, coupled with the advancement of human health risk assessment approaches, ensures that the public can continue to benefit from the advantages that pesticide use affords with the assurance that the proper use of such prod­ ucts will not result in undue harm.

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Hattis, D., Jacobs, A., Jacobson-Kram, D., Lewandowski, T., Liteplo, R., Pelkonen, O., Rice, J., Somers, D., Turturro, A., West, W., and Olin, S. (2007). Issues in the design and interpretation of chronic toxi­ city and carcinogenicity studies in rodents: approaches to dose selec­ tion. Crit. Rev. Toxicol. 37, 729–837. Ross, J. F. (2000). ECOs, FOBs, and UFOs: making sense of observa­ tional data. Toxicol. Pathol. 28, 132–136. Ross, J. F., Mattsson, J. L., and Fix, A. S. (1998). Expanded clinical observations in toxicity studies: historical perspectives and contem­ porary issues. Regul. Toxicol. Pharmacol. 28, 17–26. Seed, J., Carney, E. W., Corley, R. A., Crofton, K. M., DeSesso, J. M., Foster, P. M. D., Kavlock, R., Kimmel, G., Klaunig, J., Meek, M. E., Preston, R. J., Slikker, W. Jr., Tabacova, S., Williams, G. M., Wiltse, J., Zoeller, R. T., Fenner-Crisp, P., and Patton, D. E. (2005). Overview: using mode of action and life stage information to evaluate the human relevance of animal toxicity data. Crit. Rev. Toxicol. 35, 663–672. Smith, G. S., Hall, R. L., and Walker, R. M. (2002). Applied clinical pathol­ ogy in preclinical toxicology testing. In “Handbook of Toxicologic Pathology” (W. M. Haschek, C. G. Rousseax, and M. A. Wallig, eds.), 2nd ed.,Vol. 1, pp. 123–156. Academic Press, San Diego. Sonich-Mullin, C., Filder, R., Wiltse, J., Baetcke, K., Dempsey, J., FennerCrisp, P., Grant, D., Hartley, M., Knaap, A., Kroese, D., Mangelsdorf, I., Meek, M. E., Rice, J. M., and Younes, M. (2001). IPCS conceptual framework for evaluating a mode of action for chemical carcinogen­ esis. Regul. Toxicol. Pharmacol. 34, 146–152. U.S. EPA (1989). Pesticide Programs, Good Laboratory Practice Standard, Final Rule (40 CFR, Part 160). Fed. Reg. 54(158), 34052–34074. U.S. EPA (1998a). “Health Effects Test Guidelines, OPPTS 871.4100: Chronic Toxicity.” Washington, DC. U.S. EPA (1998b). “Health Effects Test Guidelines, OPPTS 871.4200: Carcinogenicity.” U.S. EPA,Washington, DC. U.S. EPA (1998c). “Health Effects Test Guidelines, OPPTS 871.4300: Com­ bined Chronic Toxicity/Carcinogenicity.” U.S. EPA, Washington, DC. U.S. EPA (2008). “Benchmark Dose Technical Guidance Document.” EPA/630/R-00/0001F. Risk Assessment Forum, Washington, DC. U.S. EPA (2002). “Determination of the Appropriate FQPA Safety Factor(s) in Tolerance Assessment,” Office of Pesticide Programs, Washington, DC. U.S. EPA (2003). “Rodent Carcinogenicity Studies: Dose Selection and Evaluation,” Health Effects Division (HED) Interim Guidance Document G2003.02. Office of Pesticide Programs, Washington, DC.

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U.S. EPA (2005a). “Guidelines for Carcinogen Risk Assessment,” EPA/630/ P-03/001B. Risk Assessment Forum, U.S. EPA, Washington, DC. U.S. EPA (2005b). “Supplemental Guidance for Assessing Susceptibility from Early-Life Exposure to Carcinogens,” EPA/630/R-03/003F. Risk Assessment Forum, U.S. EPA, Washington, DC. U.S. EPA (2006). “Chemicals Evaluated for Carcinogenic Potential,” Science Information Management Branch, Health Effects Division, Office of Pesticide Programs, U.S. EPA, Washington, DC. Velasquez, S. F., Schoeny, R., Cogliano, V. J., and Rice, G. E. (1995). Cancer risk assessment: historical perspectives, current issues, and future directions. In “Toxicology and Risk Assessment: Principles, Methods and Applications” (A. M. Fan and L. W. Chang, eds.), pp. 219–244. Informa Healthcare, New York. VKM (2008). “Combined Toxic Effects of Multiple Chemical Exposures. Opinion of the Scientific Steering Committee of the Norwegian Scientific Committee for Food Safety.” VHM, Oslo, Norway. WHO (1978). “World Health Organization International Programme on Chemical Safety, Environmental Health Criteria 6, Principles and Methods for Evaluating the Toxicity of Chemicals.” WHO, Geneva. WHO (1987). “World Health Organization International Programme on Chemical Safety, Environmental Health Criteria 70, Principles for the Safety Assessment of Food Additives and Contaminants in Food.” WHO, Geneva. WHO (1990). “World Health Organization International Programme on Chemical Safety, Environmental Health Criteria 104, Principles for the Toxicological Assessment of Pesticide Residues in Food.” WHO, Geneva. WHO (1993). “World Health Organization International Programme on Chemical Safety, Environmental Health Criteria 144, Principles of Evaluating Chemical Effect on the Aged Population.” WHO, Geneva. WHO (1999). “World Health Organization International Programme on Chemical Safety, Environmental Health Criteria 210, Principles for the Assessment of Risks to Human Health from Exposure to Chemicals.” WHO, Geneva. Williams, G. M., Iatropoulos, M. J., and Enzmann, H. G. (2008). Principles of testing for carcinogenic activity. In “Principles and Methods of Toxicology” (A. W. Hayes, ed.), 5th ed., pp. 1265–1316. CRC Press, Boca Raton, FL.

Chapter 15

Immunotoxicity of Pesticides Kathleen M. Brundage and John B. Barnett West Virginia University School of Medicine, Morgantown, West Virginia

15.1  Introduction 15.1.1  Immune System The immune system is a complex system consisting of multiple cell types and anatomical sites. It is responsible for detecting and eliminating foreign invaders. The cells that are part of the immune response include B lymphocytes, T lymphocytes, macrophages, dendritic cells, neutrophils, and natural killer (NK) cells. These cells must work together to effectively clear a foreign pathogen. The cells that make up the immune system have unique roles during an immune response. For example, B lymphocytes are responsible for activating T lymphocytes and producing antibodies specific for the foreign invader (LeBien and Tedder, 2008). T lymphocytes are separated into three populations, T helper cells, cytotoxic T cells (CTLs), and regulatory T cells, each with a unique role in the immune response (Heinonen and Perreault, 2008). Upon activation, T helper cells produce cytokines that assist in fully activating B lymphocytes to produce antibodies of different isotypes (IgM, IgG1–4, IgE, and IgA) and assist CTLs in their activation. CTLs upon activation by their cognate antigen kill virus- and bacteria-infected cells. Regulatory T cells are responsible for turning the immune response off once the threat has been eliminated. Macrophages have several roles in the immune response including activating T cells by presenting antigen, producing cytokines to recruit other immune cells to the site, as well as phagocytizing and killing bacteria (Mosser and Edwards, 2008). Dendritic cells are responsible for processing and presenting antigens to T lymphocytes thereby activating the T lymphocytes (Bousso, 2008). Neutrophils are usually the first to

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respond to an infection, phagocytizing and killing bacteria (Kantari et al., 2008). NK cells kill other cells in the body that express abnormal surface proteins due to a defect in the cell that can be caused by a virus or bacterial infection of the cell (Arrenberg et al., 2009). As stated previously, in order to clear an infection from the body a coordinated interaction of many if not all these cell types is required. The different cells of the immune system develop and interact at many unique anatomical sites including the primary lymphoid organs (bone marrow and thymus) and secondary lymphoid organs [spleen, gut associated lymphoid tissues (GALT), and lymph nodes]. Most of the cells of the immune system are generated in the bone marrow. The hematopoietic stem cells in the bone marrow expand and differentiate into the different immune cell types (i.e., B lymphocytes, macrophages, dendritic cells, etc.) in response to environmental signals including soluble growth factors and contact with bone marrow stromal cells. Mature, functional T lymphocytes are generated in the thymus. The thymus is seeded by immature Thy-1 bearing cells that migrate from the bone marrow. In the thymus these Thy-1 cells undergo a differentiation and selection process to expressing T cell receptors (TCRs) with unique antigen specificities. Only those thymocytes that survive the positive and negative selection process and are thus not self-reactive are allowed to exit to the periphery. It is in the secondary lymphoid organs such as the spleen, GALT, and lymph nodes where the cells of the immune system become activated and respond to invasion by a foreign pathogen. There are two types of immune responses, innate and adaptive. The innate immune response is the first response to invasion by a foreign pathogen (usually within seconds

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or minutes of infection) and is considered an antigen independent response that does not induce immunological memory. It consists of physical barriers such as the skin and mucosal membranes along with neutrophils, macrophages, and NK cells. These cells secrete inflammatory mediators including cytokines, chemokines, reactive oxygen, and nitrogen species which can kill the invading pathogen as well as attract other cells to the site to help fight the infection. Unlike the innate immune response, the adaptive immune response is an antigen-dependent response that induces immunological memory and has a lag time of a few days. There are two arms of the adaptive immune system, one is cell-mediated immunity the other is humoral immunity. Cell-mediated immunity involves killing of viral- or bacterial-infected cells by CTLs. Humoral immunity revolves around antibodies specific for the pathogen generated by B lymphocytes and helper T lymphocytes that secrete cytokines to assist the B lymphocytes. In a humoral immune response the antibodies generated eliminate the foreign pathogen by formation of immune complexes, complement fixation, opsonization, and antibody-directed cellular cytotoxity. The immune system is a complex system that has to be tightly regulated in order to function without having a detrimental effect on the body. It must differentiate foreign pathogens from self and shut down an immune response once the pathogen has been cleared. Anything that disrupts this delicate balance can have a long-term effect on the health and survival of the organism.

15.1.2  Pesticides Pesticides are a diverse group of chemicals that by definition are designed to destroy unwanted pests. They have been used for centuries throughout the world both outside and inside the home to kill insects, rodents, weeds, and fungi. When pesticides are used properly, they have proven to be very beneficial to man, increasing crop yields and preventing the spread of disease. Unfortunately, there can be unwanted consequences to pesticide use due to overapplication and misapplication. These unwanted consequences include unintended and in many cases detrimental effects on nontarget species including humans. For most of us exposure to pesticides occurs at relatively low doses for short periods of time. Usually, at low doses pesticides do not cause any permanent harm to adult humans. However, two groups of individuals are at a greater risk, agricultural workers and children. Agricultural workers are particularly vulnerable due to the amount of pesticide they are exposed to and the length of the exposure. In the literature there are examples of an association between occupational pesticide exposure and reproductive problems, neurologic dysfunction, and changes to the immune response (Colosio et al., 1999; Frazier, 2007;

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Kamel and Hoppin, 2004; Lawrence, 2007). A second vulnerable population is children, who are vulnerable because they are still developing. There are a number of reviews that discuss in detail what is known about pesticide exposure early in life and their detrimental effects on developing children (Infante-Rivard and Weichenthal, 2007; Weselak et al., 2007; Wigle et al., 2007). The conclusion of the authors of these reviews is that pesticides can alter the immune, neuronal, and reproductive systems. However, more studies are needed with better exposure estimations and identification so that more definitive correlations can be made. Normally pesticide exposure occurs either at work (occupational exposure) or at home. Occupational exposure is associated with the manufacturing, packaging, or application of pesticides. The exposure usually occurs either through the skin (dermal) or as a result of inhalation. At home, the most common route of pesticide exposure is through the ingestion of contaminated food and/or water. Particularly in rural agricultural areas, a second relatively common route of exposure is inhalation of dust containing the pesticides. Due to the pervasive use of pesticides, we are all exposed to them on a daily basis whether we work with them directly or not. When it comes to the immune system there are four potential outcomes to pesticide exposure. The first outcome, and the most common, is no alteration to the immune system as a result of pesticide exposure. The second potential outcome is an increase in activation of the immune system with the potential to develop into an autoimmune disease. A third possible outcome of pesticide exposure is a decrease in the immune system activity resulting in immunosuppression. The final potential outcome is the development of hypersensitivity. In the literature there are both in vivo and in vitro data that demonstrate that a particular pesticide can induce alterations to the immune system. Depending on the pesticide, these alterations can include a decrease in neutrophil and macrophage function, a decrease in the number of thymocytes, a decrease/increase in mitogen-induced proliferation, a decrease in antibodydependent cell cytotoxicity, and a decrease/increase in cytokine secretion (Beach and Whalen, 2006; Colosio et al., 1999; Hong et al., 2004; Lawrence, 2007; Nagayama et al., 2007). In some instances the changes to the immune system were associated with an increase in upper respiratory infections, sinusitis, and bronchitis. In the sections that follow, examples of pesticides that alter the immune system are discussed in detail.

15.2  Carbamates This group includes some of the most heavily used pesticides in the United States and around the world such as carbaryl, mancozeb, ziram, and propoxure. They control insects,

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weeds, and fungi and are used on such crops as pecans, apples, citrus fruit, and soybeans, as well as on lawns and forests. The insecticide members of this group work by interfering with acetylcholinesterase activity (Mutero et al., 1994). Some of the compounds in this group are considered to be quite harmful to humans and other nontarget species.

15.2.1  Rodent Studies A number of studies have examined the effect of different carbamate pesticides on the rodent’s immune system. The fumigant sodium methyldithiocarbamate (SMD) has been used to control Dutch elm disease. Unfortunately, it has also been demonstrated in rodent studies to be immunotoxic. Specifically, SMD has been demonstrated to inhibit production of several cytokines by peritoneal macrophages, including the proinflammatory cytokines IL-12, IL-1, IL-18, and IFN-, while increasing production of IL-10 in mice (Pruett et al., 2005). This shift in cytokine production to a more TH2-type response (elevated IL-10, decreased IL-12) could explain the exacerbation of asthma in individuals exposed to SMD. SMD induces its effect in macrophages by inhibiting MAP kinases and subsequent activation of the transcription factor AP-1 (Pruett et al., 2005). Its major metabolite, methylisothiocyanate (MITC), has been demonstrated to have similar effects in mice at doses as low as 17 mg/kg which is a level that agricultural workers using SMD can be exposed to (Pruett et al., 2005). Both MITC and SMD have been demonstrated to decrease thymus weight and cellularity along with altering leukocyte populations in the blood (Keil et al., 1996). However, unlike SMD, MITC does not affect NK cell activity (Keil et al., 1996). Other carbamate compounds have also been demonstrated to alter peritoneal macrophages. Specifically, mancozeb-exposed mice have peritoneal macrophages that when stimulated with LPS and IFN- ex vivo produced more nitrous oxide (NO) and TNF- (Chung and Pyo, 2005). But when peritoneal macrophages were exposed to mancozeb in vitro and stimulated with LPS and IFN-, there was a decrease in NO and TNF- productions (Chung and Pyo, 2005). These data suggest that the metabolite of mancozeb may enhance the ability of macrophages to respond to stimuli and kill bacteria while the parent compound inhibits these functions. Another carbamate pesticide is carbofuran. Carbofuran has also been demonstrated to inhibit cytokine production in male C57BL/6 mice. Specifically, IFN- production by macrophages and T lymphocytes was inhibited (Jeon et al., 2001). In addition, T lymphocytes from these mice secreted less IL-2 and normal levels of IL-4 (Jeon et al., 2001). Several studies have determined that carbaryl, a wide-spectrum insecticide, can be immunotoxic. In vitro, carbaryl was demonstrated to inhibit NO production by macrophages and IL-2–dependent T cell proliferation (Casale et al., 1993; Hong et al., 2004).

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Mechanistically, carbaryl has been demonstrated to inhibit LPS-induced IFN- production and activation of NFB (Igarashi et al., 2006; Ohnishi et al., 2008). In studies using a rat model, the route of exposure has been demonstrated to be important when analyzing carbaryl immunotoxicity. Specifically, inhalation of carbaryl suppresses the humoral immune response and oral exposure enhances the allergic response in the lungs (Dong et al., 1998; Ladics et al., 1994). In mice, the route of exposure determines the effect aminocarb has on the immune system. Bernier et al. (1995) demonstrated that oral and dermal route exposure stimulated the humoral immune response while the intraperitoneal (i.p.) exposure (an unlikely route of exposure in humans) decreased the response and inhalation did not affect the response. Other carbamate pesticides such as propoxur have been shown to also inhibit the humoral immune response. In one 28-day study using male Wistar rats, oral propoxur exposure decreased plaque-forming cells (PFCs) at high doses (Siroki et al., 2001). In the same study propoxur did not affect the delayed-type hypersensitivity (DTH) response in these animals (Siroki et al., 2001). In other studies, propoxur has been demonstrated to decrease humoral immune responses in mice but at dose levels 10 greater than the allowed daily intake (ADI) limit of 0.02 mg/kg/day (Hassan et al., 2004). Other evidence that carbamates inhibit the humoral immune response comes from studies with cupravit and previcur, both of which suppressed the primary and memory humoral immune response in female mice at 300 and 1000 ppm, respectively (Elsabbagh and El-tawil, 2001). Aldicarb is another carbamate pesticide that alters the humoral immune response but only under specific conditions. In one 28-day exposure study of C57BL/6 mice, 1 and 10 ppb aldicarb suppressed the PFC response and increased T cell activation (Hajoui et al., 1992). Interestingly, when mice were exposure for 90 days to the same concentrations there was no demonstrated effect on the immune system (Hajoui et al., 1992). The data from this study suggest that over time the immune system compensates for the changes induced by the aldicarb. In another study, Thomas et al. (1990) determined that there was no alteration to the immune system in female B6C3F1 exposed for 34 days to 1, 10, or 100 ppb aldicarb via their drinking water. Whether or not the lack of effect in the Thomas study was due to the mice being exposed for a longer period of time (28 vs. 34 days) or due to the use of different mouse strains in the two studies remains to be determined. However, a third study using C3 H mice demonstrated that a single i.p. injection of 1000 ppb aldicarb resulted in an inhibition of macrophage function with no effect on the T cell response as measured in a mixed lymphocyte reaction (Dean et al., 1990). Together, these data suggest that aldicarb has the potential to be toxic to the immune system of mice depending on the length of exposure, route of exposure, and mouse strain.

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15.2.2  Nonrodent Studies There have been two studies that examined the effect of carbamate pesticides on chickens because of their use to prevent insect infestation of chicken flocks. In one study, the immune response of chickens fed 20 ppm carbonyl [the no-observedeffect level (NOEL)] for 3 months was significantly suppressed (Singh et al., 2007). In this study, when compared to controls the phagocytic ability of the exposed bird’s macrophages was reduced as well as the mitogen-induced proliferation of their lymphocytes and the DTH response to tuberculin (Singh et al., 2007). In another study using the carbamate pesticide, carbendazim demonstrated that the humoral immune response of chickens was reduced at the NOEL dose (Singhal et al., 2003). Together these studies indicate that even at low levels, exposure to carbamate pesticides can be toxic to the immune system of the nontarget species chicken.

15.2.3  Human Studies Evidence from human studies indicates that some carbamate pesticides can be immunotoxic. In one study using human NK cells as little as a 1-h exposure to 2.5 m of the fungicide ziram decreased the ability of human NK cells to lyse target cells and this decrease lasted up to 6 days after exposure (Taylor et al., 2005). If exposure was performed for a longer period of time, as little as 125 nm ziram resulted in a decrease in NK cytolytic activity (Wilson et al., 2004). As a side note, ziram in addition to being a fungicide is also an additive in rubber products including latex gloves. A second study analyzed Italian vineyard workers who used the carbamate pesticide mancozeb. In this study, mancozeb exposure was correlated with a decrease in CD25 cells in the blood and a decrease in TNF- secretion by LPS stimulated macrophages (Corsini et al., 2005). In this study, the investigators also demonstrated an increase in B lymphocytes and PMA/Io (phorbol myristate and ionomycin) -induced proliferation but no changes in NK cell function or serum immunoglobulin levels (Corsini et al., 2005). A second study by Colosio et al. (2007) reported that Italian vineyard workers exposed to mancozeb had similar levels of CD4 lymphocytes, NK cells, IgA, and IgM in their serum as individuals in the control group. The authors of these two studies suggested that differences between the two studies could be explained by a difference in exposure pattern due to variation in application of mancozeb, which is known to be quite varied (Colosio et al., 2007). The data from the human studies provide some evidence that carbamate pesticides such as mancozeb and ziram are toxic to the human immune system, but further studies are needed.

15.3  Organochlorines Another large class of pesticides that have been used extensively since World War II as insecticides, fungicides, and

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herbicides is the organochlorines. This class includes such chemicals as DDT, dieldrin, and 3,4-dichloropropionanilide (DCPA or propanil). DDT was one of the most extensively used insecticides in the world. Unfortunately, DDT, like chlordane and heptachlor (other organochlorines), persists in the environment and is not degraded by any naturally occurring biological processes; as a result, these pesticides have become a major environmental problem. In addition, these compounds are known to have detrimental effects on nontarget species and as a result, DDT, chlordane, dieldrin, and heptachlor have been banned in developed nations. In this section, only those organochlorine pesticides that are still currently being used are discussed.

15.3.1  Rodent Studies Lindane (-hexachlorocyclohexane) is an insecticide that is used in warehouses to prevent infestation and as a fumigant on seeds. Several studies indicate that lindane can alter the immune response. In one study, lindane induced a prolonged induction of superoxide in thymocytes from C57BL/6 mice (Olgun et al., 2006). In another study, lindane was demonstrated to inhibit both a primary and a secondary antibody response to sheep red blood cells (SRBCs) with a greater effect on the secondary response in white albino mice (Banerjee et al., 1996). Oral exposure to lindane for 24 weeks resulted in a dose-dependent biphasic change to the immune response (Meera et al., 1993). In the beginning (4 weeks of exposure) the immune response was enhanced followed by an inhibition of the response after 24 weeks of exposure (Meera et al., 1993). These changes correlated with an increase in radiolabeled calcium intake at the beginning of the response and a decrease in uptake of calcium after longer exposure (Meera et al., 1993). In utero exposure of Swiss albino mice to lindane had different effects depending on the dose. Those mice exposed to 10 mg lindane/kilogram body weight (kg bd wt) had an increased DTH response to SRBC, increased proliferative response by their spleen cells to LPS stimulation, and an increased antibody response to SRBCs (Das et al., 1990). However, when mice were exposed to a higher dose of lindane (100 mg/kg bd wt), there was no effect on the DTH response, LPS-induced proliferation, or antibody response to SRBCs (Das et al., 1990). Finally, in weanling rats exposed for 5 weeks to lindane a decreased antibody response was observed after vaccination with typhoid vaccine (Dewan et al., 1980). Overall, the data from the rodent studies indicate that there is clear evidence that lindane is toxic to the immune system. The changes that lindane induces to the immune system is highly dependent on the exposure dose, age at time of exposure, length of exposure, and route of exposure. Another organochlorine pesticide that has been studied and determined to be immunotoxic is methoxychlor, a known endocrine disruptor. In Sprague–Dawley rats

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exposed in utero (starting at day 7 of gestation) and lactationally [up to postnatal and day (pnd) 64] to methoxychlor (100 or 1000 ppm), changes in their immune response were observed in both males and females including an increase in the antibody response and NK cell activity (White et al., 2005). In an earlier study, methoxychlor was demonstrated to have a greater effect on the offspring than the mother, with the offspring having an increase in NK cell and CD8 cell numbers and a decrease in CD4 ���� CD82 thymocytes while mothers had a decrease in the percentage of CD4 CD82 thymocytes only (Guo et al., 2002). In two studies, male and female rats had different responses to methoxychlor exposure (Chapin et al., 1997; Guo et al., 2002). Only male rats that were exposed to 50 or 150 mg/kg bd wt methoxychlor during the last week of gestation through pnd 42 had a decrease in their antibody response SRBCs (Chapin et al., 1997). The data from these studies suggest that early in life exposure to methoxychlor has the potential to alter the immune system. Another organochlorine pesticide, DCPA (propanil, 3,4-dichloroproprionanilide), has been demonstrated to be immunotoxic both in vivo and in vitro (Salazar et al., 2008). This herbicide is used to control broadleaf weeds and grasses. In a mouse animal model, DCPA has been demonstrated to alter both the innate and the adaptive immune responses. DCPA has been demonstrated to induce transient thymic atrophy and a decrease in pre-B lymphocytes and IgM B lymphocytes in the bone marrow (Cuff et al., 1996; de la Rosa et al., 2003, 2005). Both mouse and human macrophage function including cytokine production, phagocytosis, and ROS production are inhibited by DCPA exposure in vitro (Frost et al., 2000; Ustyugova et al., 2007; Xie et al., 1997b). In addition, NK function is inhibited by DCPA exposure (Barnett et al., 1992). In vitro, cytotoxic T cell responses are unaffected during initial DCPA exposure but secondary responses are dramatically inhibited even if the DCPA was removed from the cultures prior to secondary stimulation (Barnett et al., 1992; Sheil et al., 2006). Cytokine production by T cells is also inhibited by DCPA exposure (Brundage et al., 2004; Zhao et al., 1995). Mechanistically, DCPA alters normal activation signaling events in T cells and macrophages by inhibiting the calcium influx that is necessary for optimal T cell and macrophage activation (Lewis et al., 2008; Xie et al., 1997a). Antibody production is also altered by DCPA exposure. In a mouse model, Salazar et al. demonstrated that antibody production to Streptococcus pneumoniae vaccination was increased in DCPA-treated mice due to endocrine disruptive properties of DCPA (Salazar et al., 2005, 2008). There are a few other studies in the literature on other organochlorine pesticides that demonstrate their potential to be immunotoxic. Using a mouse macrophage cell line, RAW 264.7, Zhao et al. (2009) demonstrated that acetofenate induced apoptosis by inducing the generation of ROS, activation of the caspase signaling cascade, and DNA damage. In (NZBxNZW)F1 female mice, the organochlorine

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pesticide chlordecone accelerated the development of systemic lupus erythematosus (Wang et al., 2007). It was demonstrated to activate splenic B cells and germinal centers while increasing bcl-2 and shp-1 gene expression (Wang et al., 2007). Endosulfan has been demonstrated to inhibit LPS-induced TNF- production by rat peritoneal macrophages in vitro (Ayub et al., 2003). Pentachlorophenol has been demonstrated to alter the immune response in male Fischer 344 rats after a 28-day oral exposure (Blakley et al., 1998). In particular in this study, T lymphocyte proliferation along with B lymphocyte proliferation was enhanced while the antibody response to SRBCs was decreased (Blakley et al., 1998). From the rodent studies described previously, it is obvious that some organochlorines still in use today can be immunotoxic to nontarget species and that these pesticides must be used carefully.

15.3.2  Nonrodent Studies A number of studies have examined the effect of organochlorines on the immune system of nontarget species that live in aquatic environments. In one study that examined the effect of lindane and pp'-DDE on the immune system of a Mediterranean area farmed fish, gilthead seabream, the investigators determined that exposure to either pesticide increased gene transcription of IL-1, TNF-, MHC class I , MHC class II, TLR9, IgML, and TCR- genes in leukocytes (Cuesta et al., 2008). In another study, lindane exposure was demonstrated to modulate the intracellular calcium levels in PBLs and phagocytes from rainbow trout (Betoulle et al., 2000). In a different study, i.p. lindane exposure of rainbow trout decreased in B lymphocytes from head kidney 1 month after exposure but the response to Yersinia ruckeri was not altered (Dunier et al., 1994, 1995). Endosulfan (10 mg/l) has been demonstrated to modulate the phagocytic responses of leukocytes from three Australian native fish (crimson-spotted rainbowfish, Murray cod, and golden perch) (Harford et al., 2005). Additional studies on organochlorines like 4,4-DDE have demonstrated an effect on the immune system of loggerhead turtles (Keller et al., 2006). There are also suggestions that organochlorines alter the immune system of Atlantic stingrays based on studies on the St. Johns River in Florida, where high levels of organochlorines have been found in the water (Gelsleichter et al., 2006). In broiler chicks, endosulfan exposure for 8 weeks has been determined to decrease total leukocytes, T lymphocytes, B lymphocytes, and the contact hypersensitivity response to 2–4-dinitrofuorobenzene (Garg et al., 2004). In addition, the bursal weight and thymus weight were reduced (Garg et al., 2004). Histological examination of the bursa and thymus showed atrophy/hypoplasia, a decrease in follicles size with fewer lymphocytes, in addition to hemorrhagic lesions (Garg et al., 2004). Young Caspian terns from Lake Huron that were demonstrated to be exposed to pp'-DDE had a decrease in T lymphocyte

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function and enhanced antibody responses (Grasman and Fox, 2001). There were measurable changes in the immune system in swine fed a diet designed to mimic the organochlorine mixture found in Arctic aboriginal populations in utero and during lactation (Bilrha et al., 2004). The swine were exposed to the organochlorine mixture throughout gestation and lactation. Plasma concentrations of organochlorines in the sows and piglets at birth were similar to the levels found in Inuit women living in Canada. At 4 months of age, the swine exposed to the organochlorine-rich diet had an increase in mitogen-induced proliferation and an increase in the proportion of CD4 CD8 cells and CD8 DR cells. At 6 months of age, these increases persisted except for the increase in CD8 DR cells. At 8 months of age, the group given the highest dose of organochlorines had a decreased antibody response to Mycoplasma hyopneumoniae. The data from the studies on nonrodent and rodent species provide clear evidence that some organochlorine pesticides are toxic to the immune system of nontarget species.

15.3.3  Human Studies Several studies indicate that organochlorine exposure in women can alter their immune system. In one study a weak association with arthritis development and organochlorine exposure has been observed in women (Lee et al., 2007). In another study of women, an increase in endometriosis and a decrease in NK cell activity and IL-1 and IL-12 production were associated with pp'-DDE levels of 310– 770 ng/g of fat (Quaranta et al., 2006). For both women and men, organochlorine exposure has been associated with an increase in periodontal disease (Lee et al., 2008). Individuals who lived near a Superfund site in Aberdeen, North Carolina, had high levels of organochlorine in their plasma, particularly pp'-DDE (Vine et al., 2000). In these individuals (with high pp'-DDE), there were lower proliferative responses to mitogens compared to individuals who lived farther away (Vine et al., 2000). In neonates, pp'-DDE exposure has been demonstrated to be associated with an increase in IL-13 in cord blood plasma and changes in cord plasma IL-4/IFN- and IL-13/IFN- ratios (Brooks et al., 2007). In an in vitro exposure study, human PBMC exposed to pp'-DDE had a significant decrease in NK cytotoxic activity and IL-1 and IL-12 production (Quaranta et al., 2006). In a study examining DCPA-exposed individuals, DCPA exposure was associated with an increase in plasma levels of IgG1 and LPS-induced IL-6 release and a decrease in PHA-induced IL-10 and IFN- release (Corsini et al., 2007). Based on data from rodent, nonrodent, and human studies there is clear evidence that some organochlorines pesticides still in use today can be immunotoxic.

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15.4  Organophosphates Like the organochlorine pesticides, this is a broad, widely used class of insecticides. One of the first organophosphate insecticides developed, malathion (also known as carbophos, maldison, and mecaptothio) is still widely used today. It is used on a variety of crops to control sucking and chewing insects, mosquitoes, flies, lice, and other household insects. Just like organochlorines, some studies examining the immunotoxic potential of organophosphates provide clear evidence of organophosphate-induced toxicity to the immune system.

15.4.1  Rodent Studies A number of studies demonstrate that malathion has the potential to alter the immune system. In one study, female SJL/J mice exposed every other day for 28 days orally with 0.018–180 mg malathion/kg bd wt had an increased (150%) primary antibody response to SRBCs compared to controls (Johnson et al., 2002). The effect was observed at the lowest dose used, which was below the human ADI for malathion (Johnson et al., 2002). In this same study the investigators did not observe an increased mitogen-induced T or B lymphocyte proliferation (Johnson et al., 2002). These results suggest that malathion directly affects the machinery of antibody synthesis, rather than simply enhancing proliferation of lymphocytes. In other studies on mice, rats, and rabbits, investigators have demonstrated that malathion exposure can decrease both humoral and cell-mediated immune responses (Banerjee et al., 1998). In a second oral exposure study, mice orally exposed to commercial malathion had an enhanced respiratory burst in peritoneal cells with an increase in mast cell degranulation and increase phagocytosis of mast cell granules by other peritoneal cells (Rodgers and Ellefson, 1992). This study suggests that malathion may increase the incidence of allergic responses. An additional study on rat peritoneal macrophages demonstrated that malathion exposure inhibited nitrite production and TNF- production by LPS stimulation (Ayub et al., 2003). IFN- production has also been demonstrated to be inhibited by malathion exposure (Ohnishi et al., 2008). The data from these studies indicate that malathion is clearly toxicity to the rodent immune system. Diazinon is an organophosphate insecticide that is used primarily in nonagricultural setting. There is evidence in the literature that diazinon can modulate the immune system. In mice exposed to 50 mg diazinon/kg bd wt (1/5 LD50) for 30 days a gradual decrease in the levels of IL-2, IL-4, IL-10, IL-12, and IFN- in spleen cell cultures stimulated with phytohemagglutinin was observed (Alluwaimi and Hussein, 2007). When diazinon was given i.p. (25 mg diazinon/kg bd wt) to C57BL/6 mice for 28 days, decreased antibody response and DTH response to SRBC

Chapter | 15  Immunotoxicity of Pesticides

were observed (Neishabouri et al., 2004). At a lower dose (2 mg diazinon/kg bd wt per day), RBC-cholinesterase levels and DTH response were inhibited (Neishabouri et al., 2004). In a different study, diazinon was given orally to mice for 45 days in their food, which was either a normal diet or one high in protein (40%) or one high in lipids (20% corn oil) (Handy et al., 2002). These mice were determined to be immunosuppressed (Handy et al., 2002). There was hyperplasia in the spleen, thymus, and lymph nodes, hypochromic red cells, and vacuolated leukocytes with abnormal nuclei (Handy et al., 2002). Interestingly, a high-protein or -lipid diet made the immunosuppression worse (Handy et al., 2002). This study indicates that diet can be a confounding factor when examining chemicalinduced immunotoxicity. There are several other studies on the immunotoxic effect of organophosphate in rats. Rats exposed to phosphamidon, an organophosphate insecticide, for 28 days had a decreased antibody response to ovalbumin, leukocyte migration, and IFN- production while increasing TNF- production (Suke et al., 2008). In rats chronically exposed to low levels of acephate, the response of their cells to LPS was impaired. Specifically, acephate-exposed rats had lower levels of IL-1, TNF- , IL-4, and iNOS in their blood and/or brain with higher levels of corticosterone and corticotrophin-releasing factor (Singh and Jiang, 2003). They also had altered B cell and CD8 T cell populations in their blood (Singh and Jiang, 2003). In neonatal rats exposed to chlorpyrifos (pnd 1–4 or pnd 11–14 at 1 mg chlorpyrifos/kg bd wt/day), there was no effect immediately after treatment, but in adulthood their T lymphocyte response to mitogen stimulation was significantly decreased (Navarro et al., 2001). Thus, there is clear evidence that organophosphate insecticides are toxic to the immune system of rodents.

15.4.2  Nonrodent Studies As with the organochlorine insecticides, there have been a number of studies of organophosphate insecticides effect on the immune response of nontarget species. For example, exposure of the marine mollusk Mytilus edulis to azamethiphos resulted in a decreased immune response as measured by the phagocytic index in these mollusks compared to controls (Canty et al., 2007). Malathion, which has been demonstrated to be immunotoxic in rodents, is also immunotoxic in other species. Japanese Madeka exposed subchronically to malathion for 14 days had a decreased antibody response to SRBCs (Beaman et al., 1999). A 21day exposure to malathion decreased the resistance of the fish to Yersinia ruckeri infection (Beaman et al., 1999). In the American lobster (Homarus americanus) a single exposure to 5ppb malathion decreased phagocytosis by their cells for up to 3 weeks after exposure (De et al., 2004).

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The insecticide chlorpyrifos, which in rats can be immuno­ suppressive, was determined to have little immunotoxicity in Australian freshwater fish (crimson-spotted rainbowfish, silver perch and golden perch, and Murray cod) (Harford et al., 2005). In broiler chickens fed 2 ppm monocrotophos for 8 weeks, the total number of leukocytes and the number of T lymphocytes were decreased compared to controls (Garg et al., 2004). These data provide some evidence that certain organophosphate insecticides are toxic to the immune system of nonmammalian species.

15.4.3  Human Studies Several studies indicate that some organophosphate insecticides can be immunotoxic to humans. In an in vitro study, low doses (1–10 g/ml) of the chlorpyrifos metabolite, chlorpyrifos-oxon, were demonstrated to enhance LPS-induced IFN- expression by human PBMCs compared to LPS alone (Duramad et al., 2006). Human bone marrow cells exposed in vitro to the metabolites of parathion and malathion, paraoxon and malaoxon, respectively, inhibited the colony formation of erythrocytes [both burst-forming units-erythroid (BFU-E) and colony-forming units-erythroid (CFU-E)] and colony-forming units-granulocytes-macrophage (CFU-GM) in a dose-dependent manner (Gallicchio et al., 1987). In pesticide applicators that have continuously applied organophosphates, one study suggested an increase in allergic reactions and increase in leukemia (Galloway and Handy, 2003). Epidemiology studies suggest that some individuals who had applied organophosphates had decreased serum IgG, others had a decrease in IgM, and in another study white blood cell counts were reported to be abnormal in some of the participants (Galloway and Handy, 2003). The problem with most of these studies is that the applicators usually had applied more that one pesticide, so to directly correlate the change in immune function to a particular pesticide was quite difficult. However, based on the rodent and limited human studies there appears to be clear evidence that organophosphates are toxic to the immune system.

15.5  Phenoxy Compounds Members of this class of compounds are well-known and widely used to kill broadleaf weeds. This class includes one of the most heavily used compounds in both rural and urban settings, 2-4 dichlorophenoxy acetic acid, better known as 2,4-D. It is used to control many broadleaf weeds and is used on many types of land, including pasture land, cropland in summer fallow, forests, hay, and corn. As an interesting side note, the toxicity of the banned defoliating compound Agent Orange, which is made up of equal parts 2,4-D and 2,4,5-trichlorophenoxyacetic acid (2,4,5-T), was linked to the dioxin contaminant generated during the production of 2,4,5-T and not 2,4-D.

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15.5.1  Rodent Studies

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Pyrethrins are naturally occurring compounds produced by some chrysanthemum plants. Pyrethroids are semisynthetic derivatives of the chrysanthemumic acids. These compounds are used extensively as insecticides, particularly against mosquitoes. They work better at lower temperature and can be easily broken down by sunlight so they are usually sprayed at night (Gammon, 2007).

proliferation of T cells and LPS-induced proliferation of B cells (Stelzer and Gordon, 1984). A commonly used member of this group is permethrin. C57BL/6 thymocytes exposed in vitro to permethrin had an increase in superoxide production that lasted for greater than 15 min (Olgun and Misra, 2006). One study that looked at the effect of dermal exposure found that C57BL/6N mice exposed to permethrin every day for 10 or 30 days or every other day for 7 or 14 days had decreased antibody responses to SRBCs after 10 days of exposure (Punareewattana et al., 2001). In addition, their macrophage chemiluminescent response was inhibited at 2 and 10 days postexposure (Punareewattana et al., 2001). In a second study using female C57BL/6 mice, a decrease in mitogen-induced T lymphocyte proliferation, a decrease in macrophage hydrogen peroxide production, and a decrease in antibody production were observed (Prater et al., 2003). There have been a number of studies in rats to examine the immunotoxic potential of pyrethroids and pyrethrins. In male and female Wistar rats exposed every other day for 30 days orally to cypermethrin and methyl parathion, the investigators found a higher level of FSH and estradiol in treated rats than controls (Liu et al., 2006). In addition, a decrease in serum IgG levels in both males and females and an increase in IgA levels in females only were observed (Liu et al., 2006). Also, neutrophils from rats exposed to the insecticides had an increase in phagocytosis compared to controls (Liu et al., 2006). In another study, a 28-day oral exposure of 4-week-old Wistar rats to cypermethrin resulted in a decreased DTH response at doses of 22.2 and 55.4 mg cypermethrin/kg bd wt (Institoris et al., 1999). Macrophages from rats exposed for 7 days to cyhalothrin (synthetic type II pyrethroids) had decreased phagocytosis indexes and nitric oxide production compared to controls (Righi and PalermoNeto, 2005). In vivo exposure to cyhalothrin for 7 days resulted in a decrease in number of macrophages and phagocytosis (Righi et al., 2009). In addition, a decrease in basal neutrophil oxidiative burst and an increase in S. aureusinduced neutrophil oxidative burst were observed (Righi et al., 2009). Rats exposed via inhalation to deltamethrin and imiprothrin for 10 days had significant decreases in their immune response including antibody response to SRBCs and mitogen-induced proliferation (Emara and Draz, 2007). In vitro exposure of rat macrophages demonstrated a similar effect except there was also an increase in spontaneous and PMA-induced hydrogen peroxide release (Righi et al., 2005). It was suggested by the authors that these results may be explained by cyhalothrin acting indirectly on the macrophages secondary to hypothalamic pituitary adrenals axis activation (Righi et al., 2005).

15.6.1  Rodent Studies

15.6.2  Nonrodent Studies

Pyrethroids and pyrethrins are potent insecticides that are thought to have little toxicity to nontarget organisms. Based on animal studies many of the pyrethroids have been demonstrated to be immunosuppressive, inhibiting mitogen-induced

There are a number of studies on nonrodents species including chicken and fish. Studies on white leghorn chicken demonstrated that chicken lymphocytes exposed to the synthetic pyrethroid deltamethrin proliferated less and had a

Throughout the 1990s and into this century, the immunotoxic and the carcinogenic potential of phenoxy herbicides have been debated (Bond and Rossbacher, 1993; Elliott, 2005; Faustini et al., 1996; Garabrant and Philbert, 2002; Miligi et al., 2006). The majority of studies have focused on 2,4-D. In one study female CD-1 mice exposed to Tordon 202C (a mixture of 2,4-D and picloram) in their drinking water for 26 days had a reduced antibody response to SRBCs (Blakley, 1997). In a second study, i.p. exposure to 2,4-D was demonstrated to decrease the humoral immune response to the T-independent type 2 antigen phophorylcholine from S. pneumoniae (Salazar et al., 2005). In a third study, a commercial formulation of 2,4-D was given to pregnant CD-1 mice and the immune system of their offspring was analyzed (Lee et al., 2001). In this study, when the offspring were 7 weeks old their immune response was analyzed. In 2,4-D exposed offspring, a decreased mitogen response, a decrease in the number of B cells and CTL cells in the periphery was observed (Lee et al., 2001). However, there was no difference in the humoral immune response made by 2,4-D exposed and control offspring (Lee et al., 2001). These data from rodent studies do provide some evidence that 2,4-D is immunotoxic, but the age of the animal at the time of exposure appears to be important in determining how the immune system and its ability to respond to foreign invaders will be affected.

15.5.2  Human Studies A recent study suggests that these compounds do not pose a danger to applicators when properly used with the appropriate protective gear (Miligi et al., 2006). One thing to note is that appropriate protective gear is not always worn when pesticides are applied. In addition, no farmer is exposed to just one pesticide and the investigators have to rely on self-reporting of use. Both of these factors can make finding a causal relationship more challenging.

15.6  Pyrethroids and Pyrethrins

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higher level of apoptosis compared to controls (Kote et al., 2006). Broiler chickens exposed for 8 weeks via their food to 20 ppm fenvalerate (synthetic pyrethroid) had a decrease in total leukocytes and T lymphocytes compared to controls (Garg et al., 2004). In addition, B cell numbers, the dermal response to DNF, and splenic indices induced by graft-versus-host disease were decreased in those chickens exposed to the fenvalerate (Garg et al., 2004). In fish, deltamethrin exposure of Ancistrus multispinis resulted in an increase in NO production by kidney macrophages compared to controls (Pimpao et al., 2008). In Indian major carp (Cirrhinus mrigala) exposed for 45 days to synthetic -permethrin, a reduced response to the bacterial challenge with Aeromonas hydrophila was observed (Nayak et al., 2004). These data provide additional evidence that pyrethroids and pyrethrins have the potential to be immunotoxic to nontarget species.

15.6.3  Human Studies Hadnagy et al. (2003) determined that alterations to the immune system could be detected early (1-3 days) after a professional pest control operation in which a pyrethroidbased insecticide was used. However, 6-12 months later these changes were no longer detectable (Hadnagy et al., 2003). In a second study that examined the immune system of people 2 years after acute pyrethroid intoxication, investigators determined that these individuals were immunosuppressed based on the fact that they had an increase in opportunistic infections such as Candida infections of the gastrointestinal tract and recurring urinary and/or respiratory tract infections (Muller-Mohnssen, 1999). In addition, many of the individuals had developed autoimmune diseases including scleroderma-like syndrome, myasthenialike syndrome, and autoimmune hemolysis (Muller-Mohnssen, 1999). An in vitro study of bifenthrin exposure of the human T cell lines Jurkat and H9 demonstrated that there was an increase in their aggregation due to upregulation of ICAM and LFA-1 expression (Hoffman et al., 2006). Based on these studies as well as the animals studies, there is some evidence to suggest that pyrethroids and pyrethrins are toxic to the immune system at least shortly after exposure. The data also indicate that the immune system has the ability over time to compensate for the initial toxicity induced by pyrethroids and pyrethrins.

15.7  Triazines This group of compounds includes atrazine, the second most heavily used herbicide in the United States. It is a common contaminant in ground and surface waters, particularly in agricultural areas. It is used to control broadleaf and grassy weeds. As briefly discussed later, there has been much debate on whether or not atrazine and triazines in general are toxic to nontarget species, particularly reptiles.

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15.7.1  Rodent Studies There have been several studies examining the immuno­ toxicity of atrazine using both rat and mouse models. In one study atrazine was demonstrated to be immunotoxic due its inhibition of dendritic cell maturation (Pinchuk et al., 2007). In another study, oral exposure of adult B6C3F1 mice to atrazine for 14 days resulted in altered cell-mediated immune responses and decreased resistance to infection (Karrow et al., 2005). In another study, C57Bl/6 mice exposed for 14 days orally to atrazine had decreased thymocyte populations, an increased number of CD8 T lymphocytes, and a decreased number of B lymphocytes (Filipov et al., 2005). In two developmental studies, one in rats and the other in mice, exposure in utero and lactationally to atrazine was demonstrated to be immunosuppressive and that suppression persisted to some degree into adulthood ( 6 months of age) (Rooney et al., 2003; Rowe et al., 2006, 2008). Exposure to another common triazine, simazine, resulted in male C57BL/6 mice with an altered immune system compared to control mice. Specifically, oral exposure for 4 days to 300 or 600 mg simazine/kg bd wt resulted in decreased IgM and IgG responses and a lower level of mitogen-induced proliferation of B and T lymphocytes (Kim et al., 2003). In addition the macrophages from these mice produced lower levels of IL-1, IL-6, and TNF- (Kim et al., 2003). In vitro studies have also demonstrated the immunotoxic potentials of atrazine. Exposure of thioglycollateelicited mouse peritoneal macrophages to atrazine resulted in a decreased poly I:C-induced antiviral activity and IFN production along with a decrease in NO and TNF- production (Kim et al., 2002). Using the murine dendritic cell (DC) line JAWSII, atrazine was demonstrated to interfere with DC maturation as exposure decreased MHC class I surface antigens, CD86 (a costimulatory molecule), CD11b and CD11c (accessory molecules), and CD14 (myeloid developmental marker) (Pinchuk et al., 2007). In addition, primary thymic DC exposed to atrazine had lower expression of MHC class I surface antigen and CD11c (Pinchuk et al., 2007). Together, the in vivo and in vitro exposures provide some evidence that triazines, particularly atrazine, have the potential to be immunotoxic.

15.7.2  Nonrodent Studies Several studies have looked at the immunotoxic potential in reptiles. In one study, adult northern leopard frogs (Rana pipiens) exposed to 21 ppb atrazine for 8 days were found to have a decreased innate immune response (Brodkin et al., 2007). Specifically, a decrease in the number of thioglycollate-elicited peritoneal cavity macrophages was observed in the atrazine-treated frogs (Brodkin et al., 2007). In addition, these macrophages had decreased phagocytic activity compared to macrophages from controls (Brodkin et al., 2007). A second study examined the

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effect of atrazine on the larvae of Arizona salamanders. The larvae exposed to atrazine at environmentally relevant doses had increase susceptibility to Ambystoma tigrinum virus (Forson and Storfer, 2006). These two studies provide clear evidence that atrazine is toxic to the immune system of reptiles, a nontarget species.

15.7.3  Human Studies Most studies on the effect of atrazine on humans have examined reproductive toxicity but several evaluated its immunotoxic potential. An in vitro study using human NK cells demonstrated that atrazine inhibited target cell killing (Whalen et al., 2003). A second study reported that atrazine-induced inhibition of NK cell killing was due to an inhibition of lytic granule release (Rowe et al., 2007). In a study of pesticide applicators (males) and females working in the area but not applying triazine pesticides, an increase in chronic bronchitis was observed along with changes to a number of immune system parameters compared to controls (Klucinski et al., 2001). These studies suggest that there is some evidence that atrazine is toxic to the immune system.

15.8  Regulations 15.8.1  Immuntoxic Guidelines Several U.S. and European agencies have issued guidelines for immunotoxicity evaluation of chemicals. In 1998, the U.S. EPA promulgated guidelines for immunotoxicity testing (EPA OPPTS 870.7800). Although the results of more routine toxicological testing were to be considered in the overall evaluation of a pesticide, not all chemicals that modify the immune system induce changes to tissue pathology (Luster et al., 1992). Therefore, these guidelines recommended initially testing the adaptive immune response by measuring antibody production to the Tdependent antigen SRBCs. This assay was chosen because it requires the cooperative interaction of many different immune cell types to induce this active immune response. If the chemical is determined to be immunosuppressive by this assay, then flow cytometry is recommended to assess the effect of the pesticide on the phenotypic distribution of the major lymphocyte populations and subpopulations in the peripheral blood or spleen. If the chemical does not induce an effect on the T-dependent antigen response, then analysis of the innate immune response via NK assay may be performed. Use of the preceding methods was formally specified in 2007 (40 CFR Parts 9, 152, 156, 159, Federal Register 72, No. 207/Friday, October 26, 2007) as part of the toxicology data (Subpart F – Toxicology) required for all chemicals regulated by the Toxic Substances Control Act and the Federal Insecticide, Fungicide and Rodenticide Act.

Hayes’ Handbook of Pesticide Toxicology

Another possible way to identify chemicals that have the potential to be immunotoxic is to use genomics. In 2006, a workshop was held by the U.S. EPA in Research Triangle Park, North Carolina, to discuss using genomic techniques to replace or supplement current immunotoxicity screening procedures, provide insight into potential mechanisms of action, and provide data suitable for immunotoxicity hazard identification or risk assessment (Luebke et al., 2006). Several data gaps were noticed, including dose–response and kinetic data for known immunotoxic compounds as well as correlating genomic alterations to observed in vivo functional changes. Even though a genomic approach to screen chemicals for immunotoxic potential holds promise, as a routine method of risk assessment this approach is still a long way off. However, determining the mechanism of action is practicable, and microarray analysis is a practical method of exploring pathway changes that lead to alterations to the immune system. The use of genomics effectively in immunotoxicity screening will require a coordinated effort by industry, academics, and government laboratories to address the data gaps, validation, quality assurance, and protocol development. In 2002, the Food and Drug Administration Center for Drug Evaluation and Research released “Guidance for Industry: Immunotoxicology Evaluation of Investigational New Drugs,” which provided relatively specific methodology to evaluate immunotoxicity. In addition, it defines a full spectrum of adverse effects including immunosuppression, autoimmunity, and allergy. In 2006, the International Conference on Harmonization of Technical Requirements for Registration of Pharmaceutical for Human Use (ICH) issued “S8 Immunotoxicity Studies for Human Pharmaceuticals,” which is based on cause for concern. Specifically, it provided recommendations on nonclinical testing approaches to identify compounds that have the potential for immunotoxicity and then provides guidance on a weight of evidence decision-making approach for immunotoxicity testing. These guidelines apply only to immunosuppression and immunoenhancement while excluding allergies and drug-specific autoimmunity.

15.8.2  Immunotoxic Testing Testing the T-dependent antigen response has become the main focus of immunotoxicology testing because this response requires the interaction of several different immune cell types and factors, including B lymphocytes, T lymphocytes, antigen presenting cells, and cytokines. Any changes that occur to any component of the response will modify the response. In 2006, a workshop was held by the Society of Toxicology to address some of the variability in the way that laboratories assay the T-dependent response (Herzyk and Holsapple, 2007). The purpose of this workshop was to collect and discuss existing data on T-dependent antigen

Chapter | 15  Immunotoxicity of Pesticides

response from many laboratories to compare the various test protocols and the assessment of immunotoxic potential. The conclusions from the workshop were that due to the robustness of the response regardless of antigen and readout assay used, a highly standardize protocol was not necessary, and that extrapolating rodent data to human risk requires further discussion and analysis even though there is a high level of similarity between rodent and human immune systems. Through the years many different immunological assays have been used to assess the effect of chemicals on the developing immune system. The lack of a set of standard immune assays to assess the developmental immunotoxicity of chemicals has made interpretation and validation difficult. In the late 1980s, immunotoxicologists, as part of the National Toxicology Program, developed a tiered approach to assessing the immunotoxicity of chemicals in adult animals (Luster et al., 1988). After careful analysis of data from more than 50 compounds the scientists determined which assays were the best predictors of immunotoxicity and changes in host immune defense responses (Luster et al., 1988, 1992, 1993). More recently, the National Institute of Environmental Health Sciences and the National Institute for Occupational Safety and Health convened a workshop in which experts in the field of developmental immunotoxicology developed a tiered approach for assaying the developmental immunotoxicity of chemicals (Luster et al., 1993). The recommended assays were separated into three groups: an initial set of screening assays, assays for validation of a correlation between the assay endpoint and functional outcomes in humans, and assays for research development (Luster et al., 1993). The initial screening assays included analysis of the primary antibody response to a T-dependent antigen, the delayed-type hypersensitivity response, complete blood count (CBC), and the weights of the thymus, spleen, and lymph node. For the analysis of antibody responses, no specific antigen or age at the time of analysis was recommended. However, the immune system of embryos and neonates is not completely functional until a few weeks after birth; thus, analysis of the primary antibody response cannot be performed prior to 6 weeks of age. A CBC has been demonstrated to be a sensitive measure of immune system development in neonates since a reduction in any immune cell population can have profound effects on the ability of the immune system to respond to a foreign pathogen. The assays recommended in the second tier were chosen to assess the functional outcome of exposure to humans. The assays recommended included phenotypic analysis, macrophage function, and NK cell activity (Luster et al., 1993). The assays in this group were chosen for several reasons. Phenotypic analysis can provide useful information on the loss or enrichment of a particular cell population as the result of chemical exposure. For example, several studies have demonstrated that chemicals such as benzo[a]pyrene and 2,37,8-tetra-chlorodibenzo-p-dioxin

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can induce profound changes to fetal liver and fetal thymus cell populations (Hardin et al., 1992; Holladay and Luster, 1996; Kamath et al., 1998). Due to the major role of macrophages in both the innate and the adaptive immune response, their ability to secrete cytokines, phagocytize, and kill bacteria is a good way to assess immune system function. The surveillance function of NK cells and their importance in eliminating cells with abnormal phenotypes in the body make assessing their function important. The final tier of assays involves assays for research development. In this tier, general and lineage-specific assays were recommended (Luster et al., 1993). In these assays, the effects of chemical exposure on the hematopoietic process are assessed. Specifically, the effects of chemicals on the differentiation and proliferation capacity of lineage-specific and nonspecific progenitor cells are measured. With these assays, the hematopoietic process can be assayed at all stages of development, including embryonic, neonatal, and adult.

Conclusion A functional immune system is very important for survival. Chemicals such as pesticides have the potential to alter the immune system, making an individual more susceptible to infection. As in other types of pesticide-induced toxicity, the route of exposure, length of exposure, dose, and timing of exposure of a pesticide must be taken into consideration when evaluating the potential of a pesticide to be immunotoxic. In addition, these factors can also influence whether a pesticide is immunosuppressive or potentiates the immune system.

Acknowledgment The authors thank Dr. Robert Luebke for his critical reading of this chapter and his helpful suggestions.

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receptor 4, alters cytokine production by mouse peritoneal macrophages, and suppresses innate immunity. Toxicol. Sci. 87, 75–85. Punareewattana, K., Smith, B. J., Blaylock, B. L., Longstreth, J., Snodgrass, H. L., Gogal, R. M. Jr., Prater, R. M., and Holladay, S. D. (2001). Topical permethrin exposure inhibits antibody production and macrophage function in C57Bl/6N mice. Food Chem. Toxicol. 39, 133–139. Quaranta, M. G., Porpora, M. G., Mattioli, B., Giordani, L., Libri, I., Ingelido, A. M., Cerenzia, P., Di, F. A., Abballe, A., De, F. E., and Viora, M. (2006). Impaired NK-cell-mediated cytotoxic activity and cytokine production in patients with endometriosis: a possible role for PCBs and DDE. Life Sci. 79, 491–498. Righi, D. A., and Palermo-Neto, J. (2005). Effects of type II pyrethroid cyhalothrin on peritoneal macrophage activity in rats. Toxicology 212, 98–106. Righi, D. A., Xavier, F. G., and Palermo-Neto, J. (2009). Effects of type II pyrethroid cyhalothrin on rat innate immunity: a flow cytometric study. Int. Immunopharmacol. 9, 148–152. Rodgers, K., and Ellefson, D. (1992). Mechanism of the modulation of murine peritoneal cell function and mast cell degranulation by low doses of malathion. Agents Actions 35, 57–63. Rooney, A. A., Matulka, R. A., and Luebke, R. W. (2003). Developmental atrazine exposure suppresses immune function in male, but not female Sprague–Dawley rats. Toxicol. Sci. 76, 366–375. Rowe, A. M., Brundage, K. M., Schafer, R., and Barnett, J. B. (2006). Immunomodulatory effects of maternal atrazine exposure on male Balb/c mice. Toxicol. Appl. Pharmacol. 214, 69–77. Rowe, A. M., Brundage, K. M., and Barnett, J. B. (2007). In vitro atrazine exposure inhibits human natural killer cell lytic granule release. Toxicol. Appl. Pharmacol. 221, 179–188. Rowe, A. M., Brundage, K. M., and Barnett, J. B. (2008). Developmental immunotoxicity of atrazine in rodents. Basic Clin. Pharmacol. Toxicol. 102, 139–145. Salazar, K. D., de la Rosa, P., Barnett, J. B., and Schafer, R. (2005). The polysaccharide antibody response after Streptococcus pneumoniae vaccination is differentially enhanced or suppressed by 3,4-dichloropropionanilide and 2,4-dichlorophenoxyacetic acid. Toxicol. Sci. 87, 123–133. Salazar, K. D., Ustyugova, I. V., Brundage, K. M., Barnett, J. B., and Schafer, R. (2008). A review of the immunotoxicity of the pesticide 3,4-dichloropropionanalide. J. Toxicol. Environ. Health B Crit. Rev. 11, 630–645. Sheil, J. M., Frankenberry, M. A., Schell, T. D., Brundage, K. M., and Barnett, J. B. (2006). Propanil exposure induces delayed but sustained abrogation of cell-mediated immunity through direct interference with cytotoxic T-lymphocyte effectors. Environ. Health Perspect. 114, 1059–1064. Singh, A. K., and Jiang, Y. (2003). Lipopolysaccharide (LPS) induced activation of the immune system in control rats and rats chronically exposed to a low level of the organothiophosphate insecticide, acephate. Toxicol. Ind. Health 19, 93–108. Singhal, L. K., Bagga, S., Kumar, R., and Chauhan, R. S. (2003). Down regulation of humoral immunity in chickens due to carbendazim. Toxicol. in Vitro 17, 687–692. Singh, B. P., Singhal, L., and Chauhan, R. S. (2007). Immunotoxicity of carbaryl in chicken. Indian J. Exp. Biol. 45, 890–895. Siroki, O., Undeger, U., Institoris, L., Nehez, M., Basaran, N., Nagymajtenyi, L., and Desi, I. (2001). A study on geno- and immunotoxicological effects of subacute propoxur and pirimicarb exposure in rats. Ecotoxicol. Environ. Saf. 50, 76–81.

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Stelzer, K. J., and Gordon, M. A. (1984). Effects of pyrethroids on lymphocyte mitogenic responsiveness. Res. Commun. Chem. Pathol. Pharmacol. 46, 137–150. Suke, S. G., Ahmed, R. S., Pathak, R., Tripathi, A. K., and Banerjee, B. D. (2008). Attenuation of phosphamidon-induced oxidative stress and immune dysfunction in rats treated with N-acetylcysteine. Braz. J. Med. Biol. Res. 41, 765–768. Taylor, T. R., Tucker, T., and Whalen, M. M. (2005). Persistent inhibition of human natural killer cell function by ziram and pentachlorophenol. Environ. Toxicol. 20, 418–424. Thomas, P., Ratajczak, H., Demetral, D., Hagen, K., and Baron, R. (1990). Aldicarb immunotoxicity: functional analysis of cell-mediated immunity and quantitation of lymphocyte subpopulations. Fundam. Appl. Toxicol. 15, 221–230. Ustyugova, I. V., Frost, L. L., Van, D. K., Brundage, K. M., Schafer, R., and Barnett, J. B. (2007). 3,4-Dichloropropionaniline suppresses normal macrophage function. Toxicol. Sci. 97, 364–374. Vine, M. F., Stein, L., Weigle, K., Schroeder, J., Degnan, D., Tse, C. K., Hanchette, C., and Backer, L. (2000). Effects on the immune system associated with living near a pesticide dump site. Environ. Health Perspect. 108, 1113–1124. Wang, F., Roberts, S. M., Butfiloski, E. J., Morel, L., and Sobel, E. S. (2007). Acceleration of autoimmunity by organochlorine pesticides: a comparison of splenic B-cell effects of chlordecone and estradiol in (NZBxNZW)F1 mice. Toxicol. Sci. 99, 141–152. Weselak, M., Arbuckle, T. E., and Foster, W. (2007). Pesticide exposures and developmental outcomes: the epidemiological evidence. J. Toxicol. Environ. Health B Crit. Rev. 10, 41–80. Whalen, M. M., Loganathan, B. G., Yamashita, N., and Saito, T. (2003). Immunomodulation of human natural killer cell cytotoxic function by triazine and carbamate pesticides. Chem. Biol. Interact. 145, 311–319. White, K. L. Jr., Germolec, D. R., Booker, C. D., Hernendez, D. M., McCay, J. A., Delclos, K. B., Newbold, R. R., Weis, C., and Guo, T. L. (2005). Dietary methoxychlor exposure modulates splenic natural killer cell activity, antibody-forming cell response and phenotypic marker expression in F0 and F1 generations of Sprague–Dawley rats. Toxicology 207, 271–281. Wigle, D. T., Arbuckle, T. E., Walker, M., Wade, M. G., Liu, S., and Krewski, D. (2007). Environmental hazards: evidence for effects on child health. J. Toxicol. Environ. Health B Crit. Rev. 10, 3–39. Wilson, S., Dzon, L., Reed, A., Pruitt, M., and Whalen, M. M. (2004). Effects of in vitro exposure to low levels of organotin and carbamate pesticides on human natural killer cell cytotoxic function. Environ. Toxicol. 19, 554–563. Xie, Y. C., Schafer, R., and Barnett, J. B. (1997a). Inhibitory effect of 3,4-dichloro-propionaniline on cytokine production by macrophages is associated with LPS-mediated signal transduction. J. Leukoc. Biol. 61, 745–752. Xie, Y. C., Schafer, R., and Barnett, J. B. (1997b). The immunomodulatory effects of the herbicide propanil on murine macrophage interleukin-6 and tumor necrosis factor-alpha production. Toxicol. Appl. Pharmacol. 145, 184–191. Zhao, M., Zhang, Y., Wang, C., Fu, Z., Liu, W., and Gan, J. (2009). Induction of macrophage apoptosis by an organochlorine insecticide acetofenate. Chem. Res. Toxicol. 22, 504–510. Zhao, W., Schafer, R., Cuff, C. F., Gandy, J., and Barnett, J. B. (1995). Changes in primary and secondary lymphoid organ T-cell subpopulations resulting from acute in vivo exposure to propanil. J. Toxicol. Environ. Health 46, 171–181.

Chapter 16

Risk Assessment for Acute, Subchronic, and Chronic Exposure to Pesticides: Endosulfan1 Marilyn Silva and Sheryl Beauvais California Environmental Protection Agency, Sacramento, California

16.1  Introduction Partly because of widespread use internationally, persistence, and the fact that endosulfan is one of the few organochlorines still registered for use, the toxicity of endosulfan has received a great deal of scrutiny. It has also received intensive examination since it has been shown to bind to the estrogen receptor of MCF-7 cells (Soto et al., 1994, 1995) and there are concerns about whether or not it can affect the male reproductive system (Sinha et al., 1995, 1997, 2001) or neurological development (Seth et al., 1986; Zaidi et al., 1985) in animal studies. Legislation in the United States has more recently concentrated on sensitive population groups, such as infants and children. The fear is that people, especially the most vulnerable population groups, will be exposed via food, water, and air to dangerous levels of endosulfan that will cause irreversible damage. Conducting a risk assessment for endosulfan will either assuage those fears or the pesticide will be more strictly regulated and exposure decreased via mitigation. Under the Food Quality Protection Act [FQPA, 1996; U.S. Environmental Protection Agency (EPA), 2001a], for example, endosulfan was assessed for its potential as an endocrine disruptor in utero or during subsequent developmental stages in animal studies. In California, the endosulfan risk assessment was conducted by the Department of Pesticide Regulation (DPR) after a data call-in under the Birth Defects Prevention Act (California Senate Bill 950, 1984) revealed low no-observed-effect levels (NOELs) in a rat inhalation, a rabbit teratology, and a chronic dog study (Silva, 2008). These low NOELs signaled the potential for unacceptable risks for human exposures, based on current product labels. While the greatest risks for ­ endosulfan 1

The interpretations expressed are the authors’ and do not necessarily reflect policies of the Department of Pesticide Regulation of the California Environmental Protection Agency. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

exposure are occupational, another concern relates to exposure via air for infants at endosulfan application sites (bystander air), in addition to diet. For instance, endosulfan has been listed as a toxic air contaminant (TAC) in California [Air Resources Board (ARB) 2009] since the occupational and public exposures exceed safe levels. This chapter describes the process, used on behalf of the California EPA, DPR, to assess the potential risks associated with the use of endosulfan in California. It is intended to serve as an example of how regulations, developed to protect people and the environment from long-term effects of pesticides, are applied in practice. The risk assessment process for noncancer endpoints involves four steps (U.S. EPA, 2004a): hazard identification, dose–response assessments, exposure estimates, and risk characterization. Data from reports submitted by the principal registrants and from open literature studies are considered for toxicity endpoint selection. After a review of the toxicological database, toxic effects are identified in laboratory animal studies, following short-term (acute), subchronic (including reproduction and developmental), and chronic exposure. Subsequently, in the dose–response analysis, the lowest dose that does not cause an effect (NOEL) is identified for all endosulfan exposure durations. In the process of pesticide registration, studies submitted by the registrant are required under the Federal Insecticide, Fungicide and Rodenticide Act (FIFRA) guidelines [as amended for FQPA (1996; U.S. EPA, 2001a)]. DPR is required to rely on FIFRA guideline studies for pesticide toxicity evaluations; however, these studies are rarely published in the open literature and subsequently are not available to the public. An advantage to FIFRA guideline studies is that they provide essential details (e.g., detailed methodology, individual data, quality assurance, good laboratory practice) that are often lacking in open literature studies. 499

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16.2  Chemical identification Endosulfan (6,7,8,9,10,10-hexachloro- 1,5,5a,6,9,9a-hexahydro-6,9-methano-2,4,3-benzodioxathiepin-3-oxide), patented in 1956 (Ware, 1994), is a pesticide belonging to the chemical family of organochlorines, and it consists of two isomers [-: 64–67%; -: 29–32%; Maier-Bode, 1968; National Research Council Canada (NRCC), 1975]. The chemical formula is C9H6Cl6O3S with a molecular weight of 406.96 g/mol, and it is a mixture that forms an active ingredient (a.i.) with a -:-­isomer ratio of 2:1 (U.S. EPA, 2002). Pure endosulfan is a colorless crystal, but technical grade is brown in color, is poorly soluble in water (solubility  0.33 mg/l, 25°C), but is readily soluble in organic solvents (U.S. EPA, 2002). It is moderately volatile to air and adsorptive onto soil particles (DPR, 2004; U.S. EPA, 2002). The vapor pressure is 3.0    106 for -endosulfan and 7.2    107 mm Hg (25°C) for -endosulfan (DPR, 2004). The corresponding Henry’s law constant is 4.9    106 and 1.2    106 atm-m3/mol (- and -endosulfan, respectively; calculated from vapor pressure and solubility; U.S. EPA, 2002). The adsorption coefficients (Koc) were estimated to be 10,600 and 13,600 cm3/g for - and -endosulfan, respectively (U.S. EPA, 2002). The primary source of endosulfan in the environment is almost exclusively from pesticide application (no known natural sources). Endosulfan is a contact and stomach insecticide for food and nonfood crops, and it is toxic to fish and other aquatic organisms (Naqvi and Vaishnavi, 1993; Suntio et al., 1988; Toledo and Jonsson, 1992). It is a broad-spectrum nonsystemic insecticide and acaricide with contact and stomach action that is used to control sucking, chewing, and boring insects on a wide variety of vegetables, fruits, grains, cotton, and tea, as well as ornamental shrubs, vines, and trees (Tomlin, 1994). Endosulfan is applied through irrigation systems (chemigation), groundboom sprayer, airblast sprayer, rights-of-way sprayer (in maintenance of landscaped areas adjacent to roads, highways, power lines, telephone lines, canals, railroads, or other similar sites), low-pressure handwand sprayer, high-pressure handwand sprayer, backpack sprayer, fixed-wing aircraft, and dip treatment for germinating seed, seedling, bare root, and other commodities (U.S. EPA, 2002).

16.3  Environmental fate Endosulfan can be found in almost all media in the ­environment worldwide. The -isomer is more volatile and dissipative, while the -isomer is more adsorptive and persistent (Fan, 2008; Rice et al., 2002; Sappington and Khan, 2007; U.S. EPA, 2002). Moderate volatility enables it to be transported as vapor and spray drift, while moderate adsorption and persistence enables it to stay in the environment for an extended period (Lyman et al., 1990).

Hayes’ Handbook of Pesticide Toxicology

Air monitoring shows that endosulfan can volatilize from water, soil, and plant surfaces for 1–11 days postapplication. It can be transported via runoff to surface water bodies or via dust dispersion to distant areas and has been detected in areas where it was not used, such as the Lake Tahoe Basin and the Sequoia National Park in California, and even in the Arctic (LeNoir et al., 1999; McConnell et al., 1998; Patton et al., 1989). Photolysis and subsurface leaching are negligible. Endosulfan degradation can be via abiotic or biotic processes in aerobic and anaerobic conditions and it occurs primarily by oxidation and hydrolysis (Raupach et al., 2001). - and -endosulfan can be oxidized via biotic metabolism to endosulfan sulfate which is of comparable toxicity to - and -isomers but it is 2 times more persistent (t1/2  100–2148 days; Fan, 2008; Sappington and Khan, 2007; U.S. EPA, 2002). t1/2 values for - and -endosulfan in diverse soils and environmental conditions ranged from 19–124 to 42–265 days respectively, and those for the combined toxic residues (-/-isomers  endosulfan sulfate) ranged from 9 months to 6 years (U.S. EPA, 2002). They all can, when in water, hydrolyze abiotically or biotically to endosulfan diol. Endosulfan diol is more hydrophilic and less toxic. Hydrolysis is favored in neutral to alkaline media. At 25°C, estimated half-lives of - and -endosulfan were 11 and 19 days at pH 7, and 4 and 6 days at pH 9, respectively. However, at pH 5, they were more than 200 days for both - and -endosulfan (Fan, 2008; German Federal Environment Agency, 2004; Martens, 1977; Sappington and Khan, 2007; U.S. EPA, 2001c). Endosulfan is translocated to roots after application to leaves and metabolized within the plant so that the sulfate form is found in the roots. Translocation from leaves to roots is more rapid under warmer greenhouse conditions as compared to ambient outside temperatures. Degradation can be via abiotic or biotic processes in aerobic and anaerobic conditions. Oxidation and hydrolysis are the main routes for endosulfan degradation. Fungi and bacteria inhabiting the soil under aerobic conditions can degrade endosulfan producing the sulfate metabolite (fungi) or the diol (bacteria) (Fan, 2008; Kennedy et al., 2001; Sappington and Khan, 2007; U.S. EPA, 2001c). Under flooded (anaerobic) conditions soil microorganism metabolism yields primarily endosulfan diol (2–18%), endosulfan sulfate (3–8%), and endosulfan hydroxyether. Hydrolysis is increased with increasing pH so that the endosulfan halflife due to hydrolysis is decreased from 150 days at pH 5.5 to 1 day at pH 8.0. It is photolysed, with a half-life of approximately 7 days, giving endosulfan diol as the main product. Endosulfan sulfate is relatively stable to photolysis (Sethunathan et al., 2002). Bioaccumulation of the -isomer, -isomer, and sulfate metabolite occurs in aquatic (mussels, fish, shrimp, algae) animals [Ernst, 1977; Hazardous Substance Data Bank (HSDB), 1999; Novak and Ahmad, 1989; NRCC,

Chapter | 16  Risk Assessment for Acute, Subchronic, and Chronic Exposure to Pesticides: Endosulfan

1975; Roberts, 1972; Schimmel et al., 1977]. Metabolism occurs in terrestrial (mosquito, snail) and aquatic wildlife (Coleman and Dolinger, 1982; El Beit et al., 1981; Fan, 2008; Martens, 1977; NRCC, 1975; Sappington and Khan, 2007). However, it rapidly decreases to undetectable levels after animals are transferred to clean water.

16.4  Mechanism of toxicity Endosulfan binds to and blocks the Cl channel linked to the -amino-butyric acid (GABAA) receptor. It does not affect the GABA recognition site and so can be termed a “noncompetitive GABA antagonist” (Abalis et al., 1986; Ffrench-Constant, 1993; Lawrence and Casida, 1984). Because GABAA receptors are the principal inhibitory neuroreceptors in the mammalian brain, the antagonism of GABAergic neurons within the central nervous system (CNS) causes generalized brain stimulation (Abalis et al., 1986; Cole and Casida, 1986; Gant et al., 1987; Ozoe and Matsumura, 1986). When GABA binds to its receptor (GABAA), the Cl ion channels are opened, leading to an influx of Cl into neurons through an electrochemical gradient. The result is hyperpolarization of the cell membrane and inhibited neuron firing. Endosulfan prevents Cl from entering neurons, thus blocking the effect of GABA binding to its GABAA receptor, resulting in uncontrolled excitation.

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tissues, despite the lipophilicity of endosulfan and its primary metabolite, endosulfan sulfate (Dorough et al., 1978). After a single gavage, endosulfan administration (2 mg/kg) in female rats was assessed for metabolism and males for effects of bile duct cannulation. With bile duct cannulation, elimination in feces was decreased by two-thirds at 48 h but elimination in urine was unchanged. If the enterohepatic recirculation was a major path, then elimination in the urine would have increased and feces would have remained relatively unchanged. This suggests that metabolites passing from the liver to the intestine via the bile in the intact rat were not suitable for reabsorption and excretion via the kidney/urine. At 48 h, oral absorption in females (urine  bile) was approximately 59.7% for - and 39.3% for -isomer, and 13% of the  radioactivity and 10.8% of the  radioactivity in feces were metabolites. At 120 h, 88% of [14C]-endosulfan and 87% of -[14C]-endosulfan had been eliminated. Chan et al. (2005) used 5 mg/kg by gavage (1–3 doses) to male Sprague–Dawley rats to show that liver, kidney, fat, GI tract, muscle, brain, heart, lung, spleen, testis, and thyroid gland had 14C activity at 1, 2, 4, and 8 h postdose. 14C-Endosulfan-derived radioactivity in blood had a distribution half-life of 31 min and a terminal elimination half-life of 193 h. Blood concentration reached a maximum (0.36 mg/l) at 2 h post-dosing after being rapidly absorbed from the GI tract (absorption rate constant  3.07/h).

16.7  Toxicology profile 16.5  Biotransformation Endosulfan modifies the antioxidant enzymes superoxide dismutase (SOD), catalase (CAT), glutathione peroxidase (GPX), and glutathione (GSH) in rat liver, lung, and erythrocytes when administered via aerosol, indicating that endosulfan contributes to oxidative stress in some tissues (Bebe and Panemangatore, 2003). Stereoselective endosulfan sulfate formation from human recombinant P450s showed that CYP2B6, CYP3A4, and CYP3A5 metabolized the -endosulfan and CYP3A4 and CYP3A5 metabolized the -isomer (Casabar et al., 2006; Hodgson and Rose, 2008; Lee et al., 2006). Endosulfan affected glutathione (GSSG), GPX, reductase (GTR), and S-transferase (GST) activities (Bayoumi et al., 2001). Aminopyrine-Ndemethylase, aniline hydroxylase, and GST activities were induced by endosulfan in hepatic and extrahepatic tissues in the rat (Agrawal et al., 1983; Den Tonkelaar and Van Esch, 1974; Narayan et al., 1984, 1990a,b; Robacker et al., 1981; Singh and Pandey, 1989) (Figure 16.1).

16.6  Pharmacokinetics The majority of endosulfan, regardless of exposure route, is excreted rapidly in feces, with virtually no retention in

California has the same data requirements for pesticide registration as recommended by the U.S. EPA under FIFRA (U.S. EPA, 1997a,b). Studies performed according to FIFRA guidelines (submitted by registrants) are relied upon for risk assessment. Species tested in FIFRA studies are those recommended by the guidelines and in some cases tests must be performed in more than one species (e.g., chronic/oncogenicity: rat, mouse, and dog; developmental toxicity: rat and rabbit; neurotoxicity: hen and rat).

16.7.1  Acute Toxicity There were no acceptable FIFRA guideline oral LD50 studies for endosulfan. In rat, the oral LD50 was higher in males than in females (48 and 9.58 mg/kg, respectively). Death and/or clinical signs of neurotoxicity were observed for 24 h post-dosing, and necropsy in those that died revealed lung, stomach, intestinal, kidney, and adrenal pathology. Surviving rats had lung and adrenal pathology (Scholz and Weigand, 1971a,b). An acute, nose-only, aerosol inhalation study was performed with endosulfan in Wistar rats at 0 (polyethylene glycol  EtOH), 0.0036, 0.0123, 0.0288, 0.0401, and 0.0658 mg/l (0.61-F only; 2.08, 4.87, 6.78, 11.13 mg/kg) for 4 h (Hollander and Weigand, 1983). The LC50 was 0.0345 mg/l

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502

Cl

Cl Cl

Cl

C

C Cl-C-Cl C

S

Cl-C-Cl

O

C

Cl

SO2

Cl-C-Cl O

C C

Cl

Cl Cl Endosulfan Sulfate

Cl

OH

Cl

C C

Cl C C

Cl

CH

Cl-C-Cl CH

CH2OH

CH

CH2OH

C

Cl-C-Cl C

Endosulfan Lactone

Cl

O

C

C H2

C

Cl

Cl

Endosulfan

Cl

O

C

Cl

O

C

C

C

O

C

O

C

Cl

C

O H2C

Cl Endosulfan Diol

Endosulfan Hydroxyether

Cl

Cl Cl

H2 C

C C Cl-C-Cl C

Cl

C

O

C H2 Endosulfan Ether

Cl Figure 16.1  Proposed metabolic pathway in rat and sheep for endosulfan (Bebe and Panemangatore, 2003; Dorough et al., 1978; Lee et al., 2006). Phase I reactions on endosulfan are performed with P450s: CYP2B6, CYP3A4, and CYP3A5. Phase II reaction is with GST. Other enzymes involved with endosulfan metabolism are antioxidants: SOD, GPX, and CAT.

(5.84 mg/kg, males) and 0.0126 mg/l (2.13 mg/kg, females). Signs of neurotoxicity and death were dose- and time-related. There was a dose-related decrease in body weight gain in males (M) to day 3 and in females (F) to day 14. The lowestobserved-effects level (LOEL) was 0.0036 mg/l (0.61 mg/kg; no NOEL). Endosulfan-induced ocular irritation in rabbits was slight and dermal irritation (mild erythema) occurred in rabbits after one 24-h exposure. Endosulfan was not a dermal sensitizer in the guinea pig (Silva, 2008; U.S. EPA, 2002).

16.7.2  Subchronic Toxicity There were subchronic oral and inhalation studies that were acceptable according to FIFRA guidelines and many in the open literature had useful information. In the dietary study, CD rats were fed endosulfan for 13 weeks at 0, 0.64, 1.92, 3.85, or 23.41 mg/kg/day (M) and 0.75, 2.26, 4.59, or 27.17 mg/kg/day (F), followed by a 4-week recovery

(Barnard et al., 1985). Clinical signs, hematology effects, enlarged kidneys, and granular pigment in kidney and liver were observed at mid-dose and greater in both sexes. Increased epididymal and absolute kidney weights and decreased water consumption and the water-to-food intake were observed. The NOEL was 1.92 mg/kg/day (M) and 2.26 mg/kg/day (F). Endosulfan was administered by aerosol (nose only) 21 times over 29 days to 4- to 6-week-old Wistar rats at 0 (air only), 0 (ethanol–polyethylene 400 [1:1]), 0.0005, 0.0010, and 0.0020 mg/l air (0.097, 0.194, and 0.387 mg/ kg/day), followed by a 29-day observation (Hollander et al., 1984). The NOEL was 0.0010 mg/l (0.194 mg/kg/day). At 0.0020 mg/l (0.387 mg/kg/day) there were clinical signs of neurotoxicity and decreased body weight gains, and food consumption was decreased. Clinical chemistry was affected (Cl, creatinine, SGOT) at 0.0020 mg/l (0.387 mg/kg/ day). Except for neurotoxicity, effects were reversed by day 29 of recovery.

Chapter | 16  Risk Assessment for Acute, Subchronic, and Chronic Exposure to Pesticides: Endosulfan

16.7.3  Chronic Toxicity and Oncogenicity There were three acceptable dietary FIFRA guidelines studies (rat chronic/oncogenicity, mouse oncogenicity, and dog chronic). There were no tumors that were treatment-related, dose-related, or otherwise different in incidence across dose groups in any of these studies. The dog study was selected for the definitive NOEL for this risk assessment. Crl:CD (SD) BR rats were fed endosulfan in the diet for 104 weeks at 0, 0.1, 0.3, 0.6, or 2.9 mg/kg/day (M) and 0.1, 0.4, 0.7, or 3.8 mg/kg/day (F) to evaluate oncogenicity and chronic effects (Ruckman et al., 1989). At the high dose, body weight gain and absolute testis weight were decreased but were within historical control range. Kidney enlargement (F), progressive glomerulonephrosis (M/F; considered to be age-related), and glomerulonephritis were increased (M). The chronic NOEL was 0.6 mg/kg/day (M), based on an increased incidence of aneurysms in blood vessels at 2.9 mg/kg/day, which affected the pancreas, mesentery, and/or liver after week 80 and 0.7 mg/kg/day (F) based on the kidney effects at 3.8 mg/kg/day. Endosulfan was fed in diet to NMRI Hoe:NMRKf (SPF71) mice at 0, 2, 6, or 18 ppm (M: 0.28, 0.84, or 2.48 mg/kg/day; F: 0.32, 0.98, or 2.8 mg/kg/day) for 24 months, with interim sacrifices at 12 and 18 months (Donaubauer, 1988; Hack et al., 1995). Males at 2.48 mg/kg/day showed a 17% decrease in body weight but body weight gain was only decreased by 5%. Mortality was increased (M/F) at the high dose. The chronic NOEL was 0.84 (M) and 0.98 (F) mg/kg/day, based on increased mortality. Endosulfan was fed to Beagle dogs (6/sex/dose) at 0, 3, 10, 30, or 30/45/60 ppm (M: 0, 0.22, 0.57, 2.09, and 2.2/3.08/3.7 mg/kg/day; F: 0.19, 0.65, 1.98, and 1.95/2.78/3.57 mg/kg/day) for 1 year (Brunk, 1989). Dogs were treated for 54 days at 2.2 mg/kg/day (M) and 1.95 mg/ kg/day (F); for 52 days at 3.08 mg/kg/day (M) and 2.78 mg/ kg/day (F); and for 19–40 days at 3.7 mg/kg/day (M) and 3.57 mg/kg/day (F). All high-dose dogs were sacrificed on days 146 to 147, due to an onset of extreme neurotoxic signs after the dose was increased to 3.7 mg/kg/day (M) and 3.57 mg/kg/day (F). Both sexes showed neurotoxicity, which developed with increasing doses at the high dose level starting 2.5–6 h after treatment. Neurological symptoms, having to do with reflexes, were noted only at termination. Decreased male body weights and food consumption occurred at mid-dose and greater. The NOEL was 0.57 mg/ kg/day (M) and 0.65 mg/kg/day (F), based on neurotoxicity.

16.7.4  Genotoxicity Flodstrom et al. (1988) used endosulfan and metabolites (endosulfan sulfate, alcohol, ether, and lactone) in vivo and in vitro to assess carcinogenic potency [ability to enhance enzyme altered foci (EAF) in rat liver], tumor promoting ability, and inhibition of intercellular communication. EAF

503

were not induced in vivo after endosulfan treatment at 1 and 5 mg/kg/day for 10 weeks. In vitro a Chinese hamster lung fibroblast (V79) metabolic cooperation assay and a scrape loading/dye transfer assay with rat liver WB epithelial cells were also performed and showed gap junction communication was inhibited by endosulfan in both assays. Hepatocyte gap junctional intercellular communication was also inhibited by endosulfan metabolites (sulfate, lactone, and ether) in vitro (Ruch et al., 1990). Two rat in vivo tumor promotion EAF studies resulted in one positive (Fransson-Steen et al., 1992) and one negative result (Fransson-Steen and Warngard, 1992). The positive study provided evidence that -endosulfan (but not - or -  -) serves as a tumor promoter in rats with twothirds hepatectomy at highly toxic doses (15 mg/kg/day) that also induce systemic and/or neurotoxic effects. In primary rat hepatocytes, -endosulfan is a more potent inhibitor of intercellular communication than -endosulfan. However, the two isomers had similar inhibitory potency in WB-Fischer 344 rat liver epithelial cells (Fransson-Steen and Warngard, 1992). For genotoxicity, numerous studies have been performed in bacteria, yeast, mammalian cells in culture, and in vivo in laboratory animals (Adams, 1978; Arnold, 1972; Bajpayee et al., 2006; Chaudhuri et al., 1999; Cifone, 1983, 1984a,b; Daniel et al., 1986; Dikshith and Datta, 1977; Dikshith et al., 1978; Dorough et al., 1978; Dubois et al., 1996; Dzwonkowska and Hubner, 1986; Fahrig, 1974; Kurinnyi et al., 1982; Lu et al., 2000; L’vova, 1984; McGregor et al., 1988; Mellano, 1984; Milone and Hirsch, 1986; Moriya et al., 1983; NTP, 1988; Pednekar et al., 1987; Quinto et al., 1981; Sharma and Gautam, 1991; Shirasu et al., 1978, 1982; Sobti et al., 1983; Usha Rani et al., 1980; Usha Rani and Reddy, 1986; Velazquez et al., 1984; Yadav et al., 1982). There is some evidence for genotoxicity with endosulfan, especially in tests for chromosomal effects, but both positive and negative results have been reported.

16.7.5  Reproductive Toxicity The rat reproduction study was selected for the subchronic exposure interval since animals were treated for two generations. Crl:COBS(CD)BR rats were fed endosulfan in diet and each parental generation was mated twice (Edwards et al., 1984). Treatment was equivalent to 0.02, 1.0, 4.99 mg/kg/day (F0 M), 0.24, 1.23, 6.18 mg/kg/day (F0 F), and 0.23, 1.18, 5.72 mg/kg/day (F1b M) and 0.26, 1.32, 6.92 mg/kg/day (F1b F). Developmental endpoints were neonatal body weights, sex ratio, pup mortality, total litter loss, resorptions (early and late), gross morphology, litter size, anomalies, malformations, and others. There were no clinical signs of neurotoxicity. F1a pups had a slight decrease in mean litter weight on postnatal day (PND) 12 (7%) and PND 21 (9.7%), and F1b litter weight was decreased 12% at the highest dose tested (HDT). Adults

504

(F0 F; F1 M/F) had a marginal decrease in body weight gain and food consumption (F1 M) and an increase in relative liver (F0 M, F0 and F1b F) and kidney weights (F0 and F1b M) at the HDT. Body weight decreases in pups at PND 12–21 are expected at the higher treatment levels because (1) pups are transitioning from nursing to eating solid food, (2) there is decreased palatability of the treated diet, and (3) there are increased doses for pups as they receive endosulfan in diet and through milk. A dose response was not observed for pituitary, ovarian, or uterine weight effects and they were considered to be incidental. The lowest of the NOELs for each generation was the critical value. The critical systemic, reproductive, and pup NOEL used by DPR was 1.18 mg/kg/day.

Hayes’ Handbook of Pesticide Toxicology

6.25, 12.5, 25, 50, or 100 mg/kg (M) and 0, 0.75, 1.5, 3, 6, or 12 mg/kg (F) followed by a 15-day observation and a neuropathological examination (Bury, 1997). The neurotoxicological screening was performed 7 days prior to treatment initiation, 8 h post-dosing (time of peak effect), and 7 and 14 days post-dosing. The systemic NOEL was 12.5 mg/kg (M) and 1.5 mg/kg (F), based on an increase in clinical signs of neurotoxicity in males at 25 mg/kg and in females at 3 mg/kg, lasting for 1 day.

16.7.7.2  Developmental Neurotoxicity

Endosulfan is a strong neurotoxin in many species, including humans, but it dose not induce delayed neurotoxicity in hens (U.S. EPA, 2002, 2007b). FIFRA guidelines required oral/dietary neurobehavioral acute and subchronic toxicity assessments, as well as a developmental neurotoxicity study. All studies described here were performed in rat.

Endosulfan was fed in diet to mated female Wistar rats at 0, 3.74, 10.8, and 29.8 mg/kg/day from GD 6 through lactation day (LD) 21 (Gilmore et al., 2006). Offspring from 23 litters at 0, 3.73, and 10.8 mg/kg/day and pups from 21 litters at 29.8 mg/kg/day were assessed neurologically up to PND 75. The motility, numbers, and morphology of sperm from male pups were evaluated. Neuropathological examinations and morphometric analyses of selected neurological tissues from the pups were performed. The mean body weight of the dams was decreased in a dose-related manner during gestation, and this decrease persisted through lactation with the mean body weight of the dams at 10.8 mg/kg/day and greater significantly lower than controls through LD 7. The mean food consumption was likewise affected for all of the treatment groups during gestation. The report stated that the decrease in food consumption, while transitional, was likely due to palatability. The mean body weights of the pups in all of the treatment groups during lactation were decreased but there was no treatment-related effect on fetal gestation time, live births, viability, or lactation indices. Preputial separation (PPS) was marginally delayed (4–5%) in males at 10.8 mg/kg/day and greater (0  44.9 days; 3.74 mg/kg/ day  44.8 days; 10.8 mg/kg/day  47.1 days; P  0.05 at 29.8 mg/kg/day  46.8 days). There were no effects on other developmental parameters, neurobehavioral parameters, and neuropathology or morphometric analyses of brain in animals evaluated. The maternal NOEL was less than 3.74 mg/kg/day, based upon lower mean body weights (5–6%) and lower food consumption (12%). While these decreases are marginal, they are dose-related and therefore considered potential adverse effects. The developmental NOEL was also below 3.74 mg/kg/day based upon the lower mean pup body weights to the lowest dose tested (LDT  8% PND 11 only). Only at 3.74 mg/kg/day did the body weights return to normal. Body weight gain for pups was also decreased on PND 11 at 3.74 mg/kg/day and greater (later reversed at LDT).

16.7.7.1  Acute Neurotoxicity

16.7.7.3  Subchronic Neurotoxicity

Endosulfan was administered by oral gavage in a single dose to fasted Wistar rats at 0 (vehicle  2% starch mucilage),

Endosulfan was fed in diet to Wistar Crl:WI[Gl  /BRL/ Han]IGS BR rats at 0, 2.11, 13.7, and 37.2 mg/kg/day (M)

16.7.6  Developmental Toxicity FIFRA guideline developmental studies are required to be performed in two species, preferably the rat and the rabbit, in anticipation of capturing potential differences in toxicity. Sprague–Dawley rats were treated with endosulfan by gavage at 0 (corn oil), 0.66, 2.0, or 6.0 mg/kg/day during gestation days (GDs) 6–19 (Fung, 1980). The maternal NOEL of 2 mg/kg/day was based on significantly decreased mean body weight change (GD 0–20), decreased absolute body weight (GD 20), and increased clinical signs such as face rubbing and lethargy at 6 mg/kg/day. The developmental NOEL was 2 mg/kg/day, based on decreased mean fetal weights (8%) and decreased length and developmental skeletal anomalies at 6.0 mg/kg/day. Mated New Zealand White rabbits (20/dose) were gavaged with endosulfan at 0 (corn oil), 0.3, 0.7, or 1.8 mg/kg/day during GD 6–28 (Nye, 1981). At 1.8 mg/kg/day an additional 6 dams were added (total  26 dams) due to an unexpectedly high mortality. The maternal NOEL was 0.7 mg/kg/day based on increased mortality (4/20 dams died; 1/day GD 7, 10, 21, and 29) and on clinical signs of neurotoxicity that occurred during treatment at 1.8 mg/kg/day. Deaths occurred at 1.8 mg/ kg/day, beginning day 7, and clinical signs of neurotoxicity began on GD 6 after the first dose, at 1.8 mg/kg/day.

16.7.7  Neurotoxicity

Chapter | 16  Risk Assessment for Acute, Subchronic, and Chronic Exposure to Pesticides: Endosulfan

and 0, 2.88, 16.6, and 45.5 mg/kg/day (F) for 13 weeks to test for neurotoxicity (Sheets et al., 2004). Increased neurotoxicity and red nasal stain and decreased food consumption (possibly due to food palatability) and plasma ChE occurred in females. Absolute and relative kidney and liver weights were increased (M/F) at mid-dose and greater. Kidneys (M/F) at all doses had an amorphous brown-toyellow pigment in the cytoplasm of the proximal convoluted tubules and epithelium. NOELs were 37.2 (M) and 16.6 mg/kg/day (F).

16.8  Hazard Identification For acute, subchronic, and chronic dermal risk characterization oral NOELs were used. This method is acceptable since the toxicity between the two routes was similar and there were no dermal studies acceptable for risk assessment (U.S. EPA, 2004a).

16.8.1  Acute Toxicity 16.8.1.1  Oral NOEL The adverse effects observed in laboratory animals with acute oral exposure to endosulfan are summarized in Table 16.1. The effects observed in the LD50 study included death, clinical signs, and liver, kidney, intestine, lung, and adrenal toxicity (LD50 mg/kg  48 M; 10 F; LC50 mg/kg 4.87 M, 1.2, F; Scholz and Weigand, 1971a,b). Clinical signs observed in an acute oral neurotoxicity study in rats occurred 4–8 h post-dosing (NOEL mg/kg  12.5 M, 1.5 F; Bury, 1997) were the same as those observed in the LD50 studies and occurred more in females than in males. Acute effects in a rabbit gavage study were observed the first day of treatment at 1.8 mg/ kg/day in the absence of fetal effects (Nye, 1981). The developmental study had the lowest acute oral NOEL (0.7 mg/kg) and was selected for calculating acute dermal and dietary margins of exposures (MOEs).

Table 16.1  Definitive Studies for Critical NOELs Used for Risk Characterization Species

Exposure

Effect

NOEL (mg/kg/day) LOEL (mg/kg/day) Referencea

Developmental toxicity, 12days

Death, clinical signs beginning the first day of treatment

0.7

1.8 HDT

Nye (1981)

Generation diet

Maternal: ↑ liver and kidney weights Pup: ↓ body weights

1.18

5.4

Edwards et al. (1984)

0.194b

0.387

Hollander et al. (1984)

2.09 M 1.98 F

Brunk (1989)

Acute oral gavage Rabbit F

Subchronic oral� ����� Rat M/F

Subchronic inhalationa Rat M/F

6 h/day, 5 days/week, Clinical signs, ↓ 21 days, nose only body weight gain, ↓ food and water consumption, clinical chemistry parameters

Chronic oral� ����� Dog M/F

1 year diet

505

Premature termination, 0.57 M clinical signs of 0.65 F neurotoxicity; ↓ body weight gain and food consumption

HDT, highest dose tested; M, male; F, female. a This NOEL (0.194 mg/kg/day) was used for the acute and subchronic exposure durations. A 10 uncertainty factor was used to extrapolate from subchronic to an estimated chronic NOEL of 0.0194 mg/kg/day  (0.194 mg/kg/day  ÷  10)

506

16.8.1.2  Acute Inhalation NOEL While the DPR selection of the subchronic NOEL (0.001 mg/l; 0.194 mg/kg/day) for the acute NOEL involved some uncertainty, the advantages to this method rather than extrapolating from the LC50 are (1) LOELs from the acute (Hollander and Weigand, 1983), range-finding, and subchronic studies (Hollander et al., 1984) were similar (0.0036, 0.0024, and 0.002 mg/l or 0.61, 0.456, and 0.387 mg/kg/day, respectively); (2) more animals treated in the subchronic (15/sex/dose subchronic vs. 5/sex/dose acute); (3) the subchronic had a 29-day recovery versus acute with a 14-day observation; and (4) the estimate is low for an acute NOEL since acute NOELs are usually higher than subchronic. All three studies were performed at the same laboratory and in the same time frame (1983). DPR used this NOEL (Table 16.1) to estimate MOEs for occupational and general population scenarios (infant and adult bystanders at application sites).

16.8.2  Subchronic Toxicity 16.8.2.1  Oral NOEL The selected study for the critical NOEL was the rat dietary reproduction study, where parental effects were observed after an exposure of 24 weeks throughout premating, mating, gestation, lactation, and weaning for two generations (two matings/generation; Edwards et al., 1984). The oral, systemic NOEL was 1.18 mg/kg/day based on increased relative liver and kidney weights, decreased food consumption, and decreased body weights. Endpoints for both the reproduction and the subchronic dietary studies (Barnard et al., 1985) were similar but the reproduction study had a lower NOEL. It was used for dermal and dietary risk characterization (Table 16.1).

16.8.2.2  Inhalation NOEL The subchronic rat inhalation study was used for the critical NOEL (0.001 mg/l; 0.194 mg/kg/day), where endosulfan was administered by aerosol (nose only; 21 days, 6 h/day  29 days recovery; Hollander et al., 1984). The NOEL was used for the seasonal occupational and bystander inhalation risk characterization (Table 16.1).

16.8.3  Chronic Toxicity

Hayes’ Handbook of Pesticide Toxicology

chronic rat NOEL (0.6 mg/kg/day) (Ruckman et al., 1989) but at 2.0 mg/kg/day mortality and neurotoxicity occurred in dogs but was tolerated in rats.

16.8.3.2  Inhalation NOEL An acceptable chronic inhalation study was not available, so the subchronic rat inhalation study (NOEL  0.001 mg/l; 0.194 mg/kg/day; Hollander et al., 1984) was used with a 10 uncertainty factor (UF) to extrapolate from subchronic to chronic (estimated NOEL  0.194�� ���� ��� ÷ ��� ��� 10 UF  0.0194 mg/kg/day; 0.0001 mg/l; U.S. EPA, 2004a). The dose is lower than the chronic oral NOEL (0.57 mg/ kg/day), is route-specific, and was used to characterize risk for occupational and bystander scenarios (Table 16.1).

16.8.4  Genotoxicity/Oncogenicity When considering the results of all available in vivo studies performed in rats and mice, there is insufficient evidence indicating endosulfan is oncogenic. There were acceptable studies with well-designed, peer-reviewed protocols performed in rat (104-week chronic/oncogenicity) and in mouse (18 month) that resulted in no indication that endosulfan is oncogenic. Endosulfan is categorized as “A4: not classifiable as a human carcinogen” by the American Conference of Governmental Industrial Hygienists (ACGIH, 2005). Endosulfan is in “Group E: evidence of noncarcinogenicity for humans” (U.S. EPA, 2007a). The Pest Management Regulatory Agency (PMRA) stated, “Endosulfan was not carcinogenic in mice or rats and was not genotoxic” (PMRA, 2007).

16.9  Exposure Assessment Seasonal, annual, and lifetime exposure estimates for occupational handlers of endosulfan in support of aerial and high-acre aerial applications, root dip applicators, reentry workers, and bystanders at application sites are summarized in Tables 16.2–16.5. Assumptions for all exposure scenarios, unless otherwise indicated, were 47.3% dermal absorption, based on a rat study (Craine, 1988), a 70-kg body weight (Thongsinthusak et al., 1993), and inhalation absorption of 100% (U.S. EPA, 2002).

16.8.3.1  Oral NOEL

16.9.1  Occupational

In dogs, neurotoxicity was the most sensitive endpoint for chronic dietary endosulfan toxicity with a NOEL of 0.57 mg/kg/day and was used to characterize dietary and dermal risk (Table 16.1). The dog appears to be slightly more sensitive than the rat for chronic effects and was therefore selected as the definitive study with the critical NOEL (0.57 mg/kg/day; Brunk, 1989). It was similar to the

16.9.1.1  Acute, Short-Term Exposures   (1 Day–1 Week) For short-term exposures, DPR estimates the highest exposure an individual may realistically experience during or following legal endosulfan uses. For this “upper bound” of daily exposure, the estimated population 95th percentile

Chapter | 16  Risk Assessment for Acute, Subchronic, and Chronic Exposure to Pesticides: Endosulfan

507

Table 16.2  Exposure and Aggregate Estimates Calculated from PHED Data (Short-Term, Seasonal, and Annual) for Workers Handling Endosulfan in Support of Aerial Applicationsa Scenarioa,b

STADDc (mg/kg/day)

SADDg (mg/kg/day)

AADDh (mg/kg/day)

Dermal

Inhalation

Aggregatel

Dermal

Inhalation

Aggregatel

Dermal

Inhalation

Aggregatel

M/L EC

0.219

0.006

0.227

0.033

0.001

0.034

0.011

0.0003

0.011

M/L WPe,j

2.32

0.309

2.63

0.348

0.037

0.38

0.116

0.012

0.128

M/L WSP

0.168

0.017

0.187

0.040

0.004

0.044

0.014

0.001

0.015

Applicator

0.786

0.004

0.79

0.157

0.001

0.158

0.053

0.0003

0.053

Flagger

0.371

0.002

0.375

0.057

0.0002

0.057

0.019

0.00005

0.019

0.450

0.013

0.463

0.112

0.004

0.116

0.028

0.0008

0.029

M/L WP

4.77

0.635

5.40

1.19

0.127

1.32

0.298

0.032

0.330

M/L WSP

0.345

0.036

0.381

0.138

0.014

0.152

0.034

0.004

0.038

Applicator

1.62

0.007

1.63

0539

0.003

0.542

0.135

0.00007

0.135

Aeriald

High-acre aerial M/L EC e

f,i,k

NOTE: Dietary contributions are all 1% for the scenarios. a All scenarios (except airblast applicator) were based on data from the Pesticide Handlers Exposure Database (PHED, 1995). Airblast applicator exposure based on data from Smith (2005). Exposure rates and exposure estimates were rounded to three significant figures. EC, emulsifiable concentrate; M/L, mixer/loader; WP, wettable powder; WSP, water soluble pack. b Handlers were assumed to wear gloves as specified on product labels, except aerial applicators (exempt from wearing gloves under California law); respirator (except M/L using a closed system); and coveralls. M/L assumed to wear chemical-resistant apron. c Short-term absorbed daily dosage (STADD) is an upper-bound estimate calculated from the short-term exposure. Application rate is maximum rate on product labels, which varied for each scenario; acres treated/day varies by scenario. Estimates were rounded to three significant figures. Calculation: STADD  [(short-term exposure)  (absorption)  (acres treated/day)  (application rate)]/(70 kg body weight). Calculation assumptions include: dermal absorption  47.3% (Craine, 1988); body weight  70 kg (Thongsinthusak et al., 1993); inhalation rate  16.7 l/min (Andrews and Patterson, 2000); inhalation absorption  100%. d STADD estimates assumed 350 acres (142 ha) treated/day (U.S. EPA, 2001d), and a maximum application rate of 2.5 lb a.i./acre (2.8 kg a.i./ha), maximum rate on tree nuts. e Data from open pouring mixing/loading used in exposure estimate. U.S. EPA (2002) would require all WP to be packaged in WSP, and non-WSP packaging is being phased out. f STADD estimates assumed 1200 acres (486 ha) treated/day (U.S. EPA, 2001a), and a maximum application rate of 1.5 lb a.i./acre (1.7 kg a.i./ha), maximum rate on cotton. Multiple flaggers assumed for large-acre applications (U.S. EPA, 2001c), and high-acre scenarios include only M/L and applicator. g Seasonal average daily dosage (SADD) is a 90% upper confidence estimate calculated from the long-term exposure rates. Dermal absorption: 47.3% (Craine, 1988). Inhalation absorption assumed to be 100%. Body weight assumed to be 70 kg (Thongsinthusak et al., 1993). Calculation: SADD  [(longterm exposure)  (absorption)  (acres treated/day)  (application rate)]/(70 kg body weight). h Annual average daily dosage (AADD)  SADD  (annual use months per year)/(12 months in a year). i Exposure estimates assumed 40 acres (16 ha) treated/day (U.S. EPA, 2001d), and a maximum application rate of 2.5 lb a.i./acre (2.8 kg AI/ha), maximum rate on tree fruits. Annual exposure estimate based on high-use period of 2 months. j Data from open pour mixing/loading used in exposure estimate. U.S. EPA (2002) would require all WP to be packaged in WSP, and non-WSP packaging is being phased out. k Exposure estimates assumed 80 acres (32 ha) treated/day (U.S. EPA, 2001d), and a maximum application rate of 1.5 lb a.i./acre (1.7 kg a.i./ha), maximum rate on cotton. Annual exposure estimate based on high-use period of 5 months. l Aggregate  occupational (dermal  inhalation)  dietary exposure: acute dietary exposure  2.06 g/kg/day based on the 95th percentile of user-day exposure for females (13 years), nursing and chronic dietary exposure  0.17 g/kg/day (%CT; mean annual consumption for females (13 years)). Values were rounded to two significant figures.

of daily exposure is used. A higher percentile is not used because the higher the percentile, the less reliably it can be estimated and the more it tends to overestimate the population value (Chaisson et al., 1999).

16.9.1.2  Seasonal (1 Week–1 Year) and Annual (1 Year) To estimate seasonal and annual exposures, the average daily exposure is of interest because over these periods of

time, a worker is expected to encounter a range of daily exposures (i.e., DPR assumes that with increased exposure duration, repeated daily exposure at the upper-bound level is unlikely). To estimate the average, DPR uses the arithmetic mean of daily exposure (Powell, 2003). In most instances, the mean daily exposure of individuals over time is not known. However, the mean daily exposure of a group of persons observed in a short-term study is believed to be the best available estimate of the mean for an individual over a longer period.

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508

Table 16.3  Short-Term and Aggregate Exposure Estimates for Nursery Root Dip Handlers Exposurea

STADDb,c (mg/kg/day) Dermal

Inhalation

Aggregatee (% dietary contribution)

Nursery root dipd M/L EC

0.00003

0.000001

0.00003 (98%)

M/L WP

0.0003

0.00004

0.003 (40%)

Applicator

41.4

0.005

41.4 (1%)

a

EC, emulsifiable concentrate; M/L, mixer/loader; WP, wettable powder. Handlers were assumed to wear gloves, respirator, and coveralls, as specified on product labels (Beauvais, 2008). b Dermal and inhalation exposure calculated from surrogate data using the Pesticide Handlers Exposure Database (PHED) database and software (PHED, 1995). Values from PHED were rounded to three significant figures. There were no seasonal or long-term exposure estimates. c Short-term absorbed daily dosage (STADD) is an upper-bound estimate calculated from the short-term exposure. Application rate is maximum rate on product labels, which varied for each scenario; acres treated/day varies by scenario. Estimates were rounded to three significant figures. Calculation: STADD  [(short-term exposure)  (absorption)  (acres treated/day)  (application rate)]/(70 kg body weight). Calculation assumptions include: dermal absorption  47.3% (Craine, 1988); body weight  70 kg (Thongsinthusak et al., 1993); inhalation absorption  100%. d STADD estimates assumed handling of 40 gal/day, containing 1.25 lb a.i./40 gal (0.15 kg a.i./40 l), for a total of 1.25 lb a.i./day (0.56 kg a.i./day). M/L estimates from PHED. Applicator dermal exposure estimates based on RAGS-E equations (U.S. EPA, 2004b). Applicator inhalation exposure estimates based on SWIMODEL (U.S. EPA, 2003), assuming a saturated endosulfan vapor concentration (Beauvais, 2008). e Aggregate  occupational  dietary exposure: acute dietary exposure  2.06 g/kg/day based on the 95th percentile of userday exposure for females (13 years); nursing and chronic dietary exposure  0.17 g/kg/day (%CT; mean annual consumption for females (13 years)).

16.9.1.3  Handler Exposure (Short-Term, Seasonal, and Annual Endosulfan is used on several crops and can be applied using most conventional application methods. Exposure estimate for handlers involved in two example methods are presented – aerial and hand dipping of nursery stock. No exposure data are available for handlers involved in endosulfan applications via either method, and alternative approaches were used to estimate exposure. Handlers involved in aerial applications include mixer/loaders (M/Ls), applicators (pilots), and flaggers (persons who mark the location for a pesticide application while the application is occurring). Root dipping involves M/Ls and applicators (persons who dip roots by hand into pesticide solution). For handlers involved in aerial applications, and with mixing/loading solutions for root dipping, exposures were estimated using the Pesticide Handler Exposure Database (PHED, 1995). PHED was created from multiple studies in which exposure was monitored during specific handler

activities, including mixing, loading, application with several types of equipment, and flagging during aerial applications. DPR and U.S. EPA both use PHED to estimate exposures when chemical- and activity-specific data are not available, and both assume that handler exposure is primarily a function of the physical parameters of handling during the mixing, loading, and application processes, rather than the chemical properties of an active ingredient. However, while U.S. EPA uses central tendency estimates from PHED, DPR approaches PHED differently (Beauvais et al., 2007). When using PHED data to estimate short-term exposure, DPR estimates the 90% upper confidence limit (UCL) on the 95th percentile; for seasonal or annual exposure, DPR uses the 90% UCL on the arithmetic mean. The UCL is used to account for some of the uncertainty inherent in using surrogate data and to increase the confidence that the exposures are not underestimated. Estimating the UCL requires knowing the mean and standard deviation (SD) of total dermal exposure; as the SD for total body exposure is not available from PHED, each UCL is approximated by assuming that total exposure is lognormally distributed across persons and has a population coefficient of variation (CV) of 100%. The method of approximation is described in Powell (2007) and uses the fact that in any lognormal distribution with a given CV, the confidence limits are constant multiples of the arithmetic mean. No information is available on the amount of exposure individual handlers get on a seasonal or annual basis. However, data from DPR’s Pesticide Use Report (PUR) show that in many parts of the state and in many crops, endosulfan use does not occur throughout the year and that at other times, relatively few applications are made (DPR, 2008). It is reasonable to assume that an individual handler is less likely to be exposed to endosulfan during these relatively low-use intervals. Thus, rather than assume that handlers are exposed throughout the year, annual use patterns are plotted based on monthly PUR data from one or more counties with the highest use. Annual exposure to endosulfan is assumed to be limited to the months when use is relatively high (defined as 5% or more of annual use each month).

16.9.1.4  Aerial and High Acre Aerial Applications Aerial applications can be made to a variety of crops, including crops grown in orchards, vineyards, and fields. The amount of a.i. handled in association with aerial applications is estimated from the maximum application rate allowed on endosulfan product labels and from the numbers of acres treated. The maximum application rate for endosulfan applied aerially is on nut crops, 2.5 lb/acre (2.8 kg a.i./ha). The number of acres treated per day was assumed to be 350 acres/day (142 ha/day), based on the default recommended by U.S. EPA (2001d). Exposure estimates for handlers involved in aerial applications assumed that M/Ls

Chapter | 16  Risk Assessment for Acute, Subchronic, and Chronic Exposure to Pesticides: Endosulfan

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Table 16.4  Estimated Exposure for Reentry Workers (Short-term, Seasonal, Annual, and Aggregate) Exposure scenarioa

STADD (mg/kg/day)c

SADD (mg/kg/day)d

AADD (mg/kg/day)e

Workerb

Aggregatei (% diet)

Workerb

Aggregatei (% diet)

Workerb

Aggregatei (% diet)

Sweet corn, hand harvestingf

0.533

0.535 (2%)

0.075

0.075 (2%)

0.001

0.006 (2%)

Grape, cane turningg

0.335

0.337 (2%)

0.141

0.041 (2%)

0.047

0.047 (2%)

0.162

0.164 (2%)

0.004

0.004 (4%)

0.002

0.002 (8%)

Lettuce, scouting

h

a

Dislodgeable foliar residue (DFR) values from Beauvais (2008). Transfer coefficient (TC) is rate of skin contact with treated surfaces. TC references: cotton scouting (Dong, 1990); peach (Dawson, 2003); ornamental plants (Klonne et al., 2000); all other crops (U.S. EPA, 2000c). c Short-term absorbed daily dosage (STADD) calculated as described in text. Exposure estimates are for dermal route, as inhalation route assumed to be insignificant. Assumptions include: exposure duration  8 h; dermal Absorption  47.3% (Craine, 1988); body weight  70 kg (Thongsinthusak et al., 1993). d Seasonal average daily dosage (SADD) is a mean estimate of absorbed dose, calculated as described in text. Exposure estimates are for dermal route, as inhalation route assumed to be insignificant. Transfer coefficients in Beauvais (2008). e Annual average daily dosage (AADD)  ADD  (annual use months per year)/(12 months in a year). f Annual exposure estimate based on high-use period of 1 months. g Annual exposure estimate based on high-use period of 4 months. h Annual exposure estimate based on high-use period of 5 months i Aggregate  occupational  dietary exposure: acute dietary exposure  2.06 g/kg/day based on the 95th percentile females (13 years); nursing and chronic dietary exposure  0.17 g/kg/day (%CT; mean annual consumption for females (13 years)). b

and flaggers wear the clothing specified on product labels: long-sleeved shirt and pants, waterproof or chemicalresistant gloves, and shoes and socks. Under California law, applicators (pilots) are not required to wear gloves during an application. Exposures were also estimated for high-acre aerial application (HAA) of endosulfan to field crops such as alfalfa, cotton, and corn. The maximum application rate for endosulfan applied to cotton is 1.5 lb/acre (1.7 kg a.i./ha). “Acres treated (per day)” in HAA was assumed to be 1200 acres/day (486 ha/day), based on the default recommended by U.S. EPA (2001d). Additional flaggers are assumed to participate in HAA, and their exposures would be similar to flaggers in applications of 350 acres/day. Table 16.7 summarizes aerial handler exposure estimates; values reported are for total exposure (dermal  inhalation). Mitigation measures proposed by U.S. EPA (2002) would require all wettable powder (WP) to be packaged in water-soluble packages (WSPs).

16.9.1.5  Root Dip Root dipping may be done for treatment of cherry, peach, and plum seedlings for peachtree borer. The dipping solution is prepared by mixing 1.25 lb (0.568 kg) a.i. in 40 gallons (151 l) of water. Assuming that this solution is prepared and used through each workday, DPR uses PHED to estimate M/L exposure. Applicators are assumed to immerse seedling roots into a container such as a bucket or vat while grasping the seedlings just above the roots and that hands were immersed in the pesticide solution or slurry. PHED lacks data for this activity. Instead, dermal and inhalation exposure are estimated separately. Applicator dermal exposure is

estimated from equations on dermal absorption of chemicals from water in the Risk Assessment Guidance for Superfund, Part E (RAGS-E; U.S. EPA, 2004b). These are based on a two-compartment model, in which the skin is assumed to be composed of two main layers, the stratum corneum and the viable epidermis, with the stratum corneum as the main barrier. The permeability coefficient of the stratum corneum to a chemical (Kp) is estimated based on physical properties of the chemical, including the molecular weight and log Kow. The model assumes that absorption of material deposited on the skin continues long after the exposure has ended. Applicator inhalation exposure is estimated from equations in SWIMODEL (U.S. EPA, 2003). SWIMODEL uses wellaccepted screening exposure assessment equations to calculate swimmers’ total exposure expressed, modified from equations used by Beech (1980). For inhalation exposure, SWIMODEL assumes 100% absorption of inhaled chemical. Exposure estimates are based on chemical intakes only; the model does not address metabolism or excretion (U.S. EPA, 2003). Table 16.3 summarizes handler exposures associated with root dipping. Due to infrequent use, seasonal and annual exposures to endosulfan are not anticipated to occur, and only short-term exposures are estimated.

16.9.1.6  Reentry Representative exposure scenarios for reentry workers were selected as described in the document provided by the DPR WHS (Beauvais, 2008). No exposure data were available for workers reentering crops treated with endosulfan. Because of this, exposures of workers reentering crops treated with endosulfan were estimated from dislodgeable

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Table 16.5  Endosulfan Exposure Estimates (Short-Term, Seasonal, Annual, and Aggregate) for Bystanders at Application Sites Siteb

Mean STADD (mg/kg/day)c

Mean SADD (mg/kg/day)d

Mean AADD (mg/kg/day)e

Inhalation

Aggregatea

Inhalation

Aggregatea

Inhalation

Aggregatea

0.0016

0.00478 (67%)

0.00056

0.00084 (33%)

0.000047

0.000327 (86%)

0.00076

0.0028 (73%)

0.00027

0.00044 (39%)

0.000022

0.000192 (89%)

Bystander – infants East stationf Bystander – adults East stationf a

Aggregate mean occupational  dietary exposure: adult acute dietary exposure  2.06 g/kg/day based on the 95th percentile user-day exposure for females (13 years); nursing adult chronic dietary exposure  0.17 g/kg/day %CT; mean annual consumption for females (13 years), nursing dietary for infant acute dietary exposure  3.18 g/kg/day (95th percentile of user-day exposure—infants, non-nursing,  1 year) and chronic infant dietary exposure  0.28 g/kg/day (%CT; mean annual consumption—infants, non-nursing,  1 year). Values were rounded to two significant figures. b Estimates based on total endosulfan concentrations from monitoring conducted in San Joaquin County (application site for bystander exposure) in 1997 (ARB, 1998, 2004). c Short-term absorbed daily dose (STADD, mg/kg/day)  (short-term concentration)  (inhalation rate). Calculation assumptions: infant inhalation rate  0.59 m3/kg/day (Layton, 1993; U.S. EPA, 1997c); adult inhalation rate  0.28 m3/kg/day (OEHHA, 2000; U.S. EPA, 1997c; Wiley et al., 1991); inhalation absorption is assumed to be 100%. d Seasonal ADD (SADD)  (long-term concentration)  (inhalation rate). Calculation assumptions as above. e Annual ADD  (seasonal ADD)  (annual use months per year)/12. Annual bystander exposure estimates based on high-use period of 1 month, as repeated applications adjacent to an individual considered unlikely for longer intervals. f East Station was the application air monitoring site with the highest endosulfan TWA concentrations. Short-term exposure estimates were multiplied by 1.67 because the application rate used in the study (1.5 lb a.i./acre, or 1.7 kg a.i./ha) was below the maximum rate allowed on apples (2.5 lb a.i./acre, or 2.8 kg a.i./ha). Seasonal and annual exposure estimates were not adjusted for differences in application rate. Parentheses indicate the percent dietary contribution for aggregate exposure to endosulfan.

foliar residue (DFR) values and from transfer coefficients (TCs) from studies with surrogate chemicals (residue transfer assumed not chemical-specific) (Beauvais, 2008). Most reentry activities are not expected to result in pesticide exposure throughout the year. Annual exposure to endosulfan is assumed to be limited to the months when use is relatively high (defined as 5% or more of annual use each month). It was assumed that scouting occurred after all applications were completed.

16.9.2  Bystanders at Application Sites Air monitoring at application sites detected endosulfan, suggesting that the public may be exposed to endosulfan in air. Individuals might be exposed to endosulfan if they are working adjacent to fields that are being treated or have recently been treated (bystander exposure). Public exposure to airborne endosulfan was estimated based on monitoring studies of endosulfan at application sites (Beauvais, 2008).

16.10  Dietary Exposure 16.10.1  DPR Residue Database and Exposure Analysis Acute and chronic dietary exposure assessments were conducted for “total endosulfan” (- and -endosulfan)

and the main metabolite, endosulfan sulfate (U.S. EPA, 1997c), since the relative toxicity of the isomers and the sulfate metabolite are similar. The majority of the raw agricultural commodity (RAC) residue data used for the 1998 DPR endosulfan dietary exposure analysis were obtained from the following sources: (1) registrant commodity field residue studies (Hinstridge, 1968); (2) DPR residue monitoring data (DPR, 1994, 1995, 1997); and (3) USDA 1996 Pesticide Data Program (PDP) monitoring (USDA, 1994c, 1995, 1998). There were extensive findings of total endosulfan residues detected on label-approved RACs in the DPR market basket surveillance program during 1993, 1994, and 1995 (DPR, 1994, 1995, 1997). The USDA has a multiresidue screen analytical program and the results for total endosulfan are reported in the PDP. The PDP program targets RACs that are likely to be significantly consumed by infants and children. Acute and chronic analyses were conducted at DPR with the Exposure-4 and Exposure-1 programs, respectively, from the Technical Assessment Systems, Inc. EX dietary exposure software (TAS, 1996a,b). The Exposure4 program estimates the distribution of user-day (consumer-day) acute exposure for the U.S. population and specific subgroups (TAS, 1996a). A user-day is any day in which at least one food from the label-approved commodities is consumed. Potential acute dietary exposures were estimated using the highest measured residue values, the 95th percentile of all values, or the minimum detection

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Table 16.6  Acute and Chronic Dietary Exposure to Anticipated Endosulfan Residues on Raw Agriculture Commodities and the Resulting Dietary Margins of Exposure Population subgroup

DPR 1998 and 2007

DPR 1998 and 2007 (g/kg/day) Acute exposure 95th percentilea,c

Chronic exposureb,c

Acute MOEd

Subchronic MOEd

Chronic MOEd

U.S. population

1.85

0.19

378

621

3001

All infants  1 years

3.08

0.22

227

536

2597

Infant nursing  1 years

1.90

0.08

367

1475

7421

Infant non-nursing,  1 years 3.18

0.28

220

421

2039

Children 1–6 years

3.30

0.41

212

288

1407

Children 7–12 years

2.09

0.29

336

407

1943

Female 13–19 years not pregnant, not nursing

1.37

0.18

511

656

3187

Female 20 years not pregnant, not nursing

1.51

0.14

462

843

4082

Females 13–50 years

1.39

0.15

504

787

3840

Female 13 years pregnant, not nursing

1.57

0.15

441

787

3846

Females 13 years nursing

2.06

0.17

340

694

3448

Males 13–19 years

1.37

0.21

513

562

2668

Males 20 years

1.38

0.15

508

787

3725

Seniors 55 years

1.65

0.14

425

843

4132

Bold indicates groups of concern for DPR dietary risk assessment. a Exposure based on 1989–1992 Continuing Survey of Food Intakes of Individuals (CSFII; USDA, 1989–1992) and residue data from DPR (DPR, 1994, 1995, 1996a,b, 1997), Food and Drug Administration (FDA, 1998), and USDA (1994c, 1995, 1996b, 1998). Acute and chronic residue files used anticipated residue values for the commodities. b %CT  Adjustments were made for % crop treated (Carr, 1998, 2006; DPR, 2006; U.S. EPA, 2000a). c“ Acute” users were consumers (95th percentile based on deterministic, point estimates for residues; Exponent, 2004a). Chronic was “per capita” (consumers  nonconsumers; Exponent, 2004b). d MOE  NOEL ÷ exposure dose. Acute NOEL  0.7 mg/kg/day (rabbit developmental toxicity; Nye, 1981); subchronic NOEL  1.18 mg/kg/day (rat reproductive toxicity; Edwards et al., 1984); chronic NOEL  0.57 mg/kg/day (chronic dog study; Brunk, 1989). Chronic dietary exposure data used to calculate subchronic MOE. MOEs based on all label approved commodities.

limit (MDL) for each commodity (TAS, 1996a; USDA, 1989–1992; U.S. EPA, 2000b). For commodities with no detected residues, a value equal to the MDL is assigned to each commodity. When the residue values were derived from monitoring programs, the assumption is that the data represent high-end residue levels in the diet. When processing data were available, residue levels for the RACs and related food forms were reduced to account for the loss of residues due to washing and other processing methods. The Exposure-1 program estimates the chronic (annualized) average exposure for all members of a designated population subgroup (TAS, 1996b). Potential chronic dietary exposures were estimated using the average measured residues of all values for each commodity. For commodities with no detected residues, a value equal to one-half (50%) the MDL is assigned to

each commodity. When the residue values are derived from monitoring programs, the assumption is that data represent annual average level in the diet. If percentage of the crop treated (%CT) data are available, the average residue was further adjusted by this factor. Endosulfan is also used seasonally in California. The TAS program does not perform a subchronic dietary analysis; therefore, potential subchronic dietary exposures were estimated using chronic exposure data (average measured residue values of all values for each commodity). Alternatively, both the acute and the chronic dietary exposures were used as a bounding range to represent seasonal exposure. There are so many food uses for endosulfan that the “chronic” exposure levels for spring, summer, and autumn seasons and all seasons are very similar (0.19, 0.19, 0.21, and 0.19 g/kg/day with %CT; Carr,

512

1998; Silva, 2008) and do not vary as one might expect if endosulfan had limited uses. The overall exposure for the subchronic scenario would probably be closer to chronic than acute because it is unlikely that a commodity would be consumed at the highest detected residues for the entire season. For a shorter duration (e.g., 1–4 weeks), exposure may likely be closer to the acute rather than the chronic values.

16.10.2  Tolerances in California The California Food Safety Act (1989) requires that each specific commodity tolerance be evaluated individually. Tolerance is the maximum legal residue concentration of a pesticide allowed on RACs and processed foods. They are established by the U.S. EPA at efficacious levels for application rate and frequency but not expected to produce deleterious health effects in humans from dietary exposure [Federal Food, Drug, and Cosmetic Act (FFDCA), Sections 408–409; U.S. EPA, 1997b]. For a pesticide that is used on numerous commodities, tolerance assessments are conducted for selected fruits and vegetables. Generally, commodities are selected based on the potential for high levels of exposure. For endosulfan, the tolerances for the following RACs were evaluated: carrot, sweet corn, lettuce, milk fat, potato, strawberry, beans, cauliflower, spinach, peas, peach, summer squash, pear, pineapple, winter squash, broccoli, apple, melon, tomato, and grape (Carr, 1998, 2006). These RACs were selected because of high consumption rates or high tolerance values.

16.10.3  Residue Adjustments: Percentage of the Crop Treated DPR use %CT with chronic dietary exposure analysis (DPR, 2006; U.S. EPA, 2000a). %CT is an estimate of the acreage under cultivation that is actually treated with a pesticide at least once (e.g., %CT  percentage of total acreage for that crop). The default assumption is that 100% of any crop is treated with a given pesticide. When quality data are available indicating less than 100% is treated, then the default need not be used. The actual %CT varies from year to year depending upon biotic and abiotic factors. Using the existing %CT data, it is reasonable to revise the 100% treated assumption downward using more realistic pesticide treatment rates and use patterns. Commodities that used residues data from field trial studies obtained from the registrants (FMC Corporation, 1967; Gowan Corporation, 1967; Hinstridge, 1968) or state and federal monitoring data in the chronic dietary exposure assessment were considered for %CT adjustments (DPR, 1994, 1995, 1996a,b, 1997; USDA, 1994a,b, 1996a,c, 1997).

Hayes’ Handbook of Pesticide Toxicology

16.10.4  Highest Measured Acute Residue Values With the acute dietary exposure analysis for endosulfan, DPR used the 95th percentile residue values. This was because DPR used a deterministic approach by selecting values (point estimates) from the commodities with the highest residues or tolerances. Since the highest residue values were selected on an individual commodity basis, DPR considered the 95th percentile to be health protective (DPR, 2009). DPR can use a probabilistic approach (e.g., Monte Carlo) at the 99.9th percentile for a refined dietary assessment. This approach, however, was not selected by DPR for endosulfan since dietary exposure estimates were acceptable without further refinement.

16.10.4.1  Acute (Daily) Exposure Potential acute dietary exposures were estimated using the highest measured residue values, the 95th percentile of all values (DPR, 2009), or the MDL for each commodity (TAS, 1996a; USDA, 1989–1992). DPR evaluated endosulfan residues in the diet for several subgroups; however, three were selected to represent those at highest risk for occupational exposure and/or exposure to the general public (Beauvais, 2008; Silva, 2008) in addition to dietary exposure (Table 16.6). The subgroups of interest are (1) “females (13), nursing,” which represented the highest dietary exposure level (2.06 g/kg/day) for adults who could also be occupationally (dermal and air) exposed to endosulfan and also exposed as swimmers in surface water (dermal and noningested dietary); (2) “infants (non-nursing, 1 year)” was selected as a subgroup potentially receiving both a high dietary (3.18 g/kg/day) and air exposure as application site bystanders; and (3) children (1–6 years) had the highest dietary exposure of all subgroups (3.30 g/ kg/day) and also had the potential to be exposed (dermal and noningested dietary) as swimmers (Beauvais, 2008; Silva, 2008). These data are in bold in Table 16.6.

16.10.4.2  Subchronic (Seasonal) and Chronic (Annual) Exposure Potential subchronic dietary exposures were estimated using the chronic exposure data (average measured residue values of all values for each commodity), since the TAS, Inc. EX (TAS, 1996b) software does not perform subchronic analyses. The same subgroups were selected for the subchronic/ chronic dietary exposure analyses as for acute (Table 16.6).

16.10.4.3  Lifetime (Oncogenic) Dietary Exposure There is no calculated oncogenic potency factor for endosulfan since endosulfan is not considered oncogenic based on studies performed in animals and it has not

Chapter | 16  Risk Assessment for Acute, Subchronic, and Chronic Exposure to Pesticides: Endosulfan

been reported as carcinogenic in humans (Silva, 2008). Therefore, no cancer risk from lifetime (chronic) dietary exposure to endosulfan or any of its degradation products was determined.

16.11  Occupational and bystander air aggregate exposure 16.11.1  Aggregate (Dermal  Inhalation   Dietary) Exposure in Occupational Scenarios The calculation for aggregate exposure for aerial and HAA scenarios is simply adding the dermal  inhalatio n  dietary exposures. The scenarios with high dietary contributions, for instance root dip percentages of 98 and 40%, generally had very low occupational exposures.

513

16.11.1.2  The Aggregate (Inhalation  Dietary) Exposure in Bystander Air For adults and infants with aggregate exposure to endosulfan as bystanders, the dietary component for STADD, SADD, and AADD is the major exposure.

16.12  Risk characterization The human health risk from exposure to endosulfan was evaluated in California as shown in the calculations provided here (U.S. EPA, 2004a). A NOEL for a (noncancer) systemic toxic effect is divided by the estimated exposure, occupational, and/or dietary, yielding an MOE. The NOELs in this risk assessment are from animal studies. The MOE is considered to be protective of human health

Table 16.7  Margins of Exposure for Workers Handling Endosulfan in Support of Aerial Applications Scenarioa

Dermal inhalation aggregatec STADD MOEb

SADD MOEb

AADD MOEb

Aeriald M/L EC

3

32

3

35

194

30

52

65

29

M/L WP

1

1

1

3

5

2

5

2

1

M/L WP/ WSP

4

11

3

30

49

18

41

19

13

1

49

1

8

194

7

11

65

9

2

97

2

21

970

20

30

388

28

M/L EC

2

15

1

11

49

9

20

24

11

M/L WP

1

1

1

1

2

1

2

1

1

M/L WP/ WSP

2

5

1

9

14

5

17

5

4

1

28

1

2

65

2

4

277

4

Applicator Flagger High-acre aerial

Applicator

e

NOTE: Dietary contributions are all 1% for the above scenarios. a EC, emulsifiable concentrate; M/L, mixer/loader; WP, wettable powder, WSP, water soluble pack. b Margin of exposure (MOE)  critical oral NOEL/exposure dosage. Dermal: critical acute oral NOEL (used for dermal MOE determination)  0.7 mg/kg, rabbit developmental study (Nye, 1981). Subchronic (seasonal) oral NOEL (used for dermal MOE)  1.18 mg/kg/day, rat reproduction study (Edwards et al., 1984). Critical chronic (annual) oral NOEL (for dermal MOE)  0.57 mg/kg/day, chronic dog study (Brunk, 1989). Inhalation: critical acute and subchronic (seasonal) inhalation NOEL was 0.194 mg/kg/day, subchronic rat inhalation study. Chronic inhalation NOEL (subchronic inhalation NOEL ÷ 10)  0.0194 mg/kg/day (Hollander et al., 1984). Exposure doses from Table 16.2; values were rounded to whole integers. c Dietary MOE contribution to aggregate estimations were: acute  340, 95th percentile for females (13 years), nursing; chronic (used also for subchronic)  3448 (females (13 years), nursing. Aggregate MOE calculation: Aggregate multiroute MOE (MOET ) 1  1 1 1   MOE dermal MOE inhal MOE diet d Exposure estimates assumed 350 acres (142 ha) treated/day (U.S. EPA, 2001d), and an application rate of 1.5 lb a.i./acre (1.7 kg a.i./ha), maximum rate on collards, cotton, grapes, lettuce, sweet corn, and tomatoes. Annual exposure estimate based on high-use period of 4 months, based on data from DPR (2006). e Exposure estimates assumed 1200 acres (486 ha) treated/day (U.S. EPA, 2001d), and a maximum application rate of 1.5 lb a.i./acre (1.7 kg a.i./ha), maximum rate on cotton. Annual exposure estimate based on high-use period of 3 months.

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if it exceeds a particular value (usually 10 if the NOEL is based on a human study, 100 for an animal study, and 1000 if the pesticide is listed as a TAC in California and the NOEL is based on an animal study). The benchmark of 100 includes UFs (10 interspecies sensitivity; 10 intraspecies variability; Dourson et al., 2002). The NOELs used for risk characterization were previously described (dermal/oral  NOELoral: 0.70, 1.18, and 0.57 mg/kg/day, for acute, subchronic, and chronic toxicity, respectively; NOELinhalation  0.194 and estimated NOEL  0.0194 mg/ kg/day for acute/subchronic and chronic). The MOE calculations are described next.

Table 16.8  Estimated Margins of Exposure for Root Dip Handlers Using Endosulfan Scenarioa

Mean STADD MOEsb Dermal

Inhalation

Aggregatec

Nursery stock dip M/L EC

23,333

194,000

335

M/L WP

2333

4850

280

Applicator

1

39

1

a

16.12.1  MOE Calculations 16.12.1.1  Single Route Exposure (Dermal/ Dietary or Inhalation) For dermal scenarios, an oral NOEL was used:



Single route margin of error NOEL (e.g.,dermal or inhalation)  Exposure dosage (route specific: dermal or inhalation)

EC, emulsifiable concentrate; M/L, mixer/loader; WP, wettable powder, Dip, nursery stock dip for treating cherry, peach and plum seedlings for peach tree borer. Handlers assumed wearing gloves, respirator, coveralls; as specified on product labels (Beauvais, 2008). NOTE: No exposure data for seasonal or annual scenarios. b Margin of exposure (MOE)  critical oral NOEL/exposure dosage. Dermal: critical acute oral NOEL (used for dermal MOE determination)  0.7 mg/kg, rabbit developmental study (Nye, 1981). Subchronic (seasonal) oNOEL (used for dermal MOE)  1.18 mg/kg/ day, rat reproduction study (Edwards et al., 1984). Critical chronic (annual) oral NOEL (for dermal MOE)  0.57 mg/kg/day, chronic dog study (Brunk, 1989). Inhalation: critical acute and subchronic (seasonal) inhalation NOEL was 0.194 mg/kg/day, subchronic rat inhalation study. Chronic inhalation NOEL (subchronic inhalation NOEL ÷ 10)  0.0194 mg/kg/day (Hollander et al., 1984). Exposure doses from Table 16.3; values were rounded to whole integers. c Dietary MOE contribution to aggregate estimations were: acute  340, 95th percentile for females (13 years), nursing; chronic (used also for subchronic)  3448 (females (13 years), nursing. Aggregate MOE calculation:

16.12.1.2  Aggregate MOE Single Route Exposure Reentry exposure was dermal (oral NOEL: 0.7, 1.18, and 0.57 mg/kg/day for acute, subchronic, and chronic, respectively). NOELs were divided by the reentry exposure plus the dietary exposure (adult [females 13 year, nursing]: acute dietary  2.06 g/kg/day; chronic dietary  0.17 g/kg/day).



Aggregate MOE (oral route) NOEL (oral)  Exposure occup  exposure diet





Aggregate multiroute MOE (MOET ) 1  1 1 1   MOE dermal MOE inhal MOE diet



16.12.2  MOE Results 16.12.2.1  MOEs and Aggregate MOEs for Single Route of Exposure Primarily the STADD MOEs were less than 100 whether from a single exposure or an aggregate single route of exposure. Inhalation MOEs for bystanders at endosulfan application sites were less than 1000 for all scenarios.

16.12.1.3  Aggregate MOE: Multiroute Exposure (Inhalation  Dermal)

16.12.2.2  Aggregate MOEs for Multiple Routes of Exposure

Aggregate multiroute (inhalation  dermal dietary) aerial, root dip, or bystander air MOEs were calculated as shown below (U.S. EPA, 2001b). This approach is used when common effects occur for different routes of exposure and the same UFs apply for each route.

More than 90% of the occupational aggregate scenarios had MOEs less than 100. Bystander aggregate scenarios were all less than 1000.



Aggregate multiroute MOE (MOE T ) 1  1 1 1   MOE dermal MOE inhal MOE diet

16.13  Issues related to sensitive populations Endosulfan, like other organochlorines such as DDT, dicofol, methoxychlor, and others, many of which have been banned in the United States, is primarily a neurotoxicant

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Table 16.9  Margins of Exposure for Reentry Workers Exposure scenarioa

STADD MOEb

SADD MOEb

Mean AADD MOEb

Occupational

Aggregatec

Occupational

Aggregatec

Occupational

Aggregatec

Sweet corn, hand harvesting

1

1

16

15

95

92

Grape, cane turning

2

2

8

8

12

12

Lettuce, scouting

4

4

295

283

285

263

a

Reentry exposure scenarios from Table 16.6. Margin of exposure (MOE)  critical oral NOEL/exposure dosage. Dermal: critical acute oral NOEL (used for dermal MOE determination)  0.7 mg/kg, rabbit developmental study (Nye, 1981). Subchronic (seasonal) oral NOEL (used for dermal MOE)  1.18 mg/kg/day, rat reproduction study (Edwards et al., 1984). Critical chronic (annual) oral NOEL (for dermal MOE)  0.57 mg/kg/day, chronic dog study (Brunk, 1989). Inhalation: critical acute and subchronic (seasonal) inhalation NOEL was 0.194 mg/kg/day, subchronic rat inhalation study. Chronic inhalation NOEL (subchronic inhalation NOEL ÷ 10)  0.0194 mg/kg/day (Hollander et al., 1984). Exposure doses from Table 16.4. Values were rounded to whole integers. c Aggregate  aggregate occupational and dietary exposure. Acute dietary exposure  2.06 g/kg/day based on the 95th percentile of user-day exposure for females (13 years), nursing and chronic dietary exposure  0.17 g/kg/d [%CT; mean annual consumption for females (13 years)]. Values were rounded to two significant figures. b

Table 16.10  Estimated Endosulfan Bystander Margins of Exposure Sitea

Mean STADDb MOEs

Mean SADDb MOEs

Mean AADDb MOEs

Inhalation

Aggregatec

Inhalation

Aggregatec

Inhalation

Aggregatec

121

78

346

296

413

343

255

146

719

595

882

702

Bystander – infants East station Bystander – adults East station a

For further description see Table 16.5. Margin of exposure (MOE)  inhalation: critical acute and subchronic (seasonal) inhalation NOEL was 0.194 mg/kg/day rat subchronic inhalation study. Chronic inhalation NOEL (subchronic NOEL ÷ 10)  0.0194 mg/kg/day (Hollander et al., 1984). Exposure doses see Table 16.5. c Dietary MOE contribution to aggregate estimations were: acute  340, 95th percentile for females (13 years), nursing; chronic (used also for subchronic)  3448 (females (13 years), nursing. Aggregate MOE  1 ÷ (1 ÷ (MOE inhal)  1 ÷ (MOE diet)). b

(Plimmer and Gammon, 2003). However, because endosulfan has weak estrogenic properties (Soto et al., 1994, 1995) and because of results obtained in studies that were unacceptable for regulatory purposes (Sinha et al., 1995, 1997, 2001), it has been labeled a developmental and reproductive toxicant and an endocrine disruptor (PANNA, 2008). The risk assessment concern, of course, is that these presumptive effects resulting from endosulfan exposure during critical developmental stages (in utero, or to infants and children) will result in endocrine disruption and subsequent neurotoxicity, developmental, or reproductive adverse effects that are irreversible. The data presented in a review by Silva and Gammon (2009) do not support the case that endosulfan is a developmental or reproductive toxicant, or an endocrine disruptor. The weight of evidence is that while some studies showed effects that raised concerns, the results of studies that included peer-reviewed protocols, both positive and negative controls (such as FIFRA guideline

s­ tudies), and had results that were reproduced did not support these concerns. Animals treated in utero, as neonates or as pups (or throughout all developmental stages), do not experience effects at doses lower than those inducing toxicity in adults. Unfortunately, unlike the FIFRA guideline studies, many of the literature reports contain inadequate numbers of replicates to demonstrate statistically significant effects. Moreover, negative and positive controls are usually inadequate in these literature studies.

16.14  Endocrine disruption and the fqpa safety factor The FQPA (FQPA, 1996; U.S. EPA, 2001a) mandated U.S. EPA to “upgrade its risk assessment process as part of the tolerance setting procedures” by requiring an explicit finding that dietary tolerances are safe for children

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(U.S. EPA, 1997a,b, 1998). An extra 10-fold safety factor (10 SF) was intended to account for pre- and postnatal developmental effects or the lack of a complete database (unless U.S. EPA determined, based on reliable data, that a different margin would be safe). In 2002, the U.S. EPA Reregistration Eligibility Decision (RED: U.S. EPA, 2002) used a 10 SF required under the FQPA (1996) due to database uncertainties for endosulfan. This resulted in an unacceptable dietary risk for children 1–6 years. The FQPA amended the Federal Food, Drug, and Cosmetic Act (FFDCA) and Federal Insecticide, Rodenticide, and Fungicide Act (FIFRA) to require greater health and environmental protection for pesticide use (U.S. EPA, 1997a,b). It specifically requires consideration for the protection of infants and children using an additional safety factor in setting an exposure standard. The U.S. EPA updated its dietary risk assessment in 2007 after the database uncertainties for developmental neurotoxicity (DNT; Gilmore et al., 2006) and subchronic neurobehavioral toxicity (Sheets et al., 2004) studies were satisfied. Subsequently, the U.S. EPA reduced the FQPA SF from 10 to 1. DPR agreed with the decision and rationale for decreasing the SF to 1.

16.15  Tolerance assessment updated by DPR for 2006 In 1998, there were 72 commodities with human consumption that had U.S. EPA endosulfan tolerances (U.S. EPA, 1999, 2001a). The U.S. EPA announced in the Federal Register (U.S. EPA, 2006a,b) the final rule to revoke, remove, modify, and establish tolerances that were named in the RED (U.S. EPA, 2002). Because there are no active registrations, 9 of these tolerances have since been revoked (artichoke, canola, mustard seed, raspberry, safflower seed, sugar beet, sugarcane, sunflower seed, and watercress; U.S. EPA, 2007b). Tolerances for 5 commodities, frequently consumed by infants and children, were revoked (beans [succulent], grape [including juice and raisin], peas [succulent], pecan, and spinach), leaving 58 remaining tolerances (U.S. EPA, 2007b). The U.S. EPA concluded that the revocation of the preceding tolerances would mitigate acute dietary exposure concerns to acceptable levels for infants and children (U.S. EPA, 2002).

16.16  Summary 16.16.1  Occupational and Bystander MOEs In aerial, high-acre aerial, root dip, reentry scenarios for all exposure intervals, the majority (90%) of dermal, inhalation, and aggregate MOEs were less than 100 and therefore endosulfan represents a potential health risk for those exposed occupationally. Infant and adult bystanders at

endosulfan application sites had inhalation and aggregate MOEs of less than 1000 (additional 10 UF due to listing as a TAC in California; TAC, 1984) and this poses an unacceptable health risk for exposure via air (ARB, 2009).

16.16.2  Dietary MOES The MOEs from anticipated endosulfan residues for acute toxicity (95th percentile) were all above 100 for all population subgroups and exposure durations. Therefore, endosulfan dietary risk was characterized as health protective.

Conclusion There is ample evidence that endosulfan can be acutely poisonous to humans through accidental and intentional exposure as documented in California by Beauvais (2008), in the United States (U.S. EPA, 2002), and throughout the world (WHO, 2006), and the effect observed is generally neurotoxicity. Endosulfan is highly toxic to wildlife such as fish. Usually fish (e.g., yellow tetra: Hyphessobrycon bifasciatus) are killed as a result of endosulfan discharge into rivers, as well as application to wetlands at recommended rates (Jonsson and Toledo, 1993). Endosulfan has been considered by some to be a persistent organic pollutant (IPEN, 2008) throughout the world because it has been detected in Europe in lakes in the Alps, Pyrenees, and Caledonian Mountains (Carrera et al., 2002). It has also been detected in multiple media throughout the Arctic (air, soil, water, plants, and amphibian; Fan, 2008; Usha and Harikrishnan, 2005). For these reasons, mitigation to protect humans and the environment is necessary.

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Chapter | 16  Risk Assessment for Acute, Subchronic, and Chronic Exposure to Pesticides: Endosulfan

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USDA (1997). “Vegetables Summary – Agricultural Chemical Usage 1996.” National Agricultural Statistics Service, U.S. Department of Agriculture, Washington, DC. USDA (1998). “Pesticide Data Program (PDP) Annual Summary Calendar Year 1996,” (E. E. Figueroa and R. L. Epstein, eds.), U.S. Department of Agriculture, Agricultural Marketing Service, Washington, DC. U.S. EPA (1997a). The Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA) and Federal Food, Drug, and Cosmetic Act (FFDCA) as Amended by the Food Quality Protection Act (FQPA) of August 3, 1996. Document no. 730L97001, March 1997. Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1997b). 1996 Food Quality Protection Act Implementation Plan. March, 1997. Office of Prevention, Pesticides and Toxic Substances (7506C), U.S. Environmental Protection Agency, Washington, D.C. http://www.epa.gov/fedrgstr. U.S. EPA (1997c). “Exposure Factors Handbook,” EPA/600/P-95/002Fa. Office of Research and Development, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1998). “Health Effects Test Guidelines,” OPPTS 870.1000. EPA 712-C-98–189, Prevention, Pesticides and Toxic Substances (7101), U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1999). “Code of Federal Regulations Title 40, section 180.182 (endosulfan).” U.S. Government Printing Office, Washington, DC. U.S. EPA (2000a). Available Information on Assessing Exposure from Pesticides in Food. A User’s Guide. Original Document June 21, 2000. Updated in 2005. U.S. Environmental Protection Agency, Office of Prevention, Pesticides and Toxic Substances, Washington, DC. http://www.epa.gov. U.S. EPA (2000b). Choosing a Percentile of Acute Dietary Exposure as a Threshold of Regulatory Concern. U.S. Environmental Protection Agency, Office of Prevention, Pesticides and Toxic Substances, Washington, DC. http://www.epa.gov. U.S. EPA (2000c). Agricultural Transfer Coefficients, Policy Number 003.1 Science Advisory Council for Exposure. Revised August 7. U.S. EPA (2001a). “The Food Quality Protection Act (FQPA) Background.” Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, D.C. http://epa.gov/oppfead1/fqpa/ backgrnd.htm. U.S. EPA (2001b). “General Principles for Performing Aggregate Exposure and Risk Assessment.” U.S. Environmental Protection Agency, Office of Pesticide Programs, Washington, DC. U.S. EPA (2001c). “Environmental Fate and Ecological Risk Assessment for the Reregistration Eligibility Decision on Endosulfan.” U.S. EPA, Office of Prevention, Pesticides, and Toxic Substances, Washington, DC. U.S. EPA (2001d). Standard Values for Daily Acres Treated in Agriculture. Policy Number 009.1, Science Advisory Council for Exposure. Revised September 25. U.S. EPA (2002). Reregistration Eligibility Decision for Endosulfan. Case 0014. Office of Prevention, Pesticides and Toxic Substances, U.S. Environmental Protection Agency, Washington, DC. http://www.epa. gov/oppsrrd1/REDs/endosulfan_red.pdf. U.S. EPA (2003). “User’s Manual: Swimmer Exposure Assessment Model (SWIMODEL) Version 3.0.” Office of Pesticide Programs, Antimicrobials Division, U.S. Environmental Protection Agency, Washington, DC. http://www.epa.gov/oppad001/swimodelusersguide. pdf. U.S. EPA (2004a). An Examination of EPA Risk Assessment Principles and Practices: Staff Paper Prepared for the U.S. Environmental

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Protection Agency by Members of the Risk Assessment Task Force. Office of the Science Advisor, U.S. EPA, Washington, DC. U.S. EPA (2004b). “Risk Assessment Guidance for Superfund (RAGS), Volume 1: Human Health Evaluation Manual (Part E, Supplemental Guidance for Dermal Risk Assessment).” http://www.epa.gov/oswer/ riskassessment/ragse/pdf/chapter3.pdf2004. U.S. EPA (2006a). Endosulfan, fenarimol, imazalil, oryzalin, sodium acifluorfen, trifluralin, and ziram; proposed tolerance actions (proposed rule). U.S. EPA, Washington, DC. In Federal Register 71(80), 24615–24627. U.S. EPA (2006b). Endosulfan, fenarimol, imazalil, oryzalin, sodium acifluorfen, trifluralin, and ziram; tolerance actions (final rule). U.S. EPA, Washington, DC. In Federal Register 71(179), 54423–54434. U.S. EPA (2007a). OPPTS Harmonized Test Guidelines Series 870 Health Effects Test Guidelines – Final Guidelines. June 21, 2007. http://www.epa.gov/opptsfrs/publications/OPPTS_Harmonized/870_ Health_Effects_Test_Guidelines/Series/. U.S. EPA (2007b). Endosulfan. Hazard Characterization and Endpoint Selection Reflecting Receipt of a Developmental Neurotoxicity Study and Subchronic Neurotoxicity Study. PC Code: 079401. DP Barcode D338576. Office of Prevention, Pesticides and Toxic Substances, U.S. EPA, Washington, DC. April 7, 2007. http://www. epa.gov/pesticides/reregistration/endosulfan. Usha, S., and Harikrishnan, V.R. (2005). IPEN Pesticide Working Group Project 2004. Endosulfan—Fact sheet and answers to common questions. IPEN Pesticide Working Group Secretariat, Kerala, India. www.thanal.org.

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Usha Rani, M. V., and Reddy, P. P. (1986). Cytogenetic effects of aldrin and endosulfan in mice. IRSC Med. Sci. 14, 1125–1126. Usha Rani, M. V., Reddi, O. S., and Reddy, P. P. (1980). Mutagenicity studies involving aldrin, endosulfan, dimethoate, phosphamidon, carbaryl and ceresan. Bull. Environm. Contam. Toxicol. 25, 277–282. Velazquez, A., Creus, A., Xamena, N., and Marcos, R. (1984). Mutagenicity of the insecticide endosulfan in Drosophila melanogaster. Mutat. Res. 136, 115–118. Ware, G.W. (1994). Insecticides. In “The Pesticide Book,” pp. 44–46. Thomson, Fresno, CA. WHO (2006). Endosulfan. In “Rotterdam Convention on the Prior Informed Consent Procedure for Certain Hazardous Chemicals and Pesticides,” International Trade Chemical Review Committee; Second meeting, Geneva, 13–17 February 2006. Wiley, J. A., Robinson, J. P., Piazza, T., Garrett, K., Cirksena, K., Cheng, Y. T., and Martin, G. (1991). “Activity Patterns of California Residents. Contract No. A6–177–33. Final Report.” Air Resources Board, Research Division, California Environmental Protection Agency, Sacramento, CA. http://www.arb.ca.gov/research/abstracts/a6-177-33.htm. Yadav, A. S., Vashishat, R. K., and Kakar, S. N. (1982). Testing of endosulfan and fenitrothion for genotoxicity in Saccharomyces cerevisiae. Mutat. Res. 105, 403–407. Zaidi, N. F., Agrawal, A. K., Anand, M., and Seth, P. K. (1985). Neonatal endosulfan neurotoxicity: behavioral and biochemical changes in rat pups. Neurobehav. Toxicol. Teratol. 7, 439–442.

Section III

Emergining Approaches in Safety Evaluation

(c) 2011 Elsevier Inc. All Rights Reserved.

Chapter 17

Genetic Polymorphism and Susceptibility to Pesticides Erin C. Peck and David L. Eaton University of Washington, Seattle, Washington

17.1  Introduction In the broadest and most literal sense, gene–environment interactions refer to phenotypic effects resulting from interactions between genes and the environment. Such interactions are often viewed in the context of disease, specifically how one’s risk for an environmentally induced disease or condition is influenced by his/her genetic makeup. The current gene–environment interaction paradigm is well illustrated by the relationship of metabolism genes and exposure to pesticides. In the case of pesticide exposure, primary health outcomes of interest include neurodegenerative disorders and cancers. It is hypothesized that an exposed individual’s ability to activate and/or detoxify pesticide compounds is associated with his/her risk for the development of pesticide-related conditions and diseases. At the intersection of gene–environment interactions are biotransformation enzymes. The most well-known site of biotransformation is the liver, where biotransformation enzymes are found in abundance. Extrahepatic tissues also contain these enzymes, though often in smaller quantities, and thus possess their own biotransformation capabilities. Biotransformation enzymes act on endogenous compounds, such as steroid hormones, as well as exogenous compounds, including pesticides and other man-made chemicals, often serving to make them more polar and more suitable for excretion. Sometimes, however, biotransformation may generate more reactive products, which may be more toxic than their parent compounds. This property is both enzyme and substrate dependent. Specific proteins and classes of proteins, such as paraoxonase (PON1), cytochromes P450 (CYPs), and glutathione S-transferases (GSTs), have been well studied with regard to their roles in the activation and deactivation of pesticide compounds. CYPs are phase I enzymes which typically catalyze oxidation reactions. Depending on the Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

pesticide substrate, this process may result in an activation or detoxification product. The insecticide parathion, for example, is metabolized by CYPs to its “active” form, which is paraoxon. Paraoxon exerts its toxicity through the inhibition of cholinesterase, which is the mode of action for many organophosphorous (OP) compounds. Human serum PON1 is known for its ability to catalyze the hydrolysis of paraoxon to p-nitrophenol and diethylphosphoric acid, which are nontoxic compounds. Like PON1, GSTs are also thought to perform the detoxification of pesticide parent compounds or their metabolites. These phase II enzymes catalyze the attachment of glutathione to the parent compound or reactive intermediate, facilitating its excretion from the body. CYPs, GSTs, and PON1 are just a few of the enzymes thought to be involved in the activation and detoxification of pesticide compounds. Polymorphisms in the genes that encode the previously described enzymes may render an individual susceptible to or protected from the harmful effects of a particular pesticide. Subsequently, the presence of these genetic variants may serve to explain the differential susceptibility of individuals to pesticide-related diseases.

17.2  Diseases with putative “gene– pesticide” interactions Pesticide exposures have been the subject of many epidemiology studies seeking to identify associations between particular chronic diseases and pesticides.

17.2.1  Pesticides, Genes, and Cancer Risk As an example, there have been dozens of studies investigating a potential association between the widely used herbicide, 2,4-D, and lymphoma and leukemia (reviewed in 525

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Garabrant and Philbert, 2002; Ibrahim et al., 1991; Miligi et al., 2006). Although positive associations between a specific pesticide and a particular type of cancer have been found in some studies, there is often a great deal of inconsistency among such studies, and one commonly invoked reason for this is genetic heterogeneity in susceptibility. Yet there is very little research to date that has actually explored “gene–environment” interactions that might modify associations between pesticide exposure and particular cancers. From a mechanistic perspective, common genetic variants in genes involved with pesticide metabolism and disposition, such as the CYPs, GSTs, PON1, epoxide hydrolase (mEH), etc., would be logical candidates for association studies in pesticide-exposed populations (Eaton, 2000). From the perspective of cancer risk, it would be most logical to focus such studies on pesticides that are mutagenic and/or have tested positive in cancer bioassays. However, fortunately there are relatively few pesticides on the market today that have such characteristics, and thus it is particularly difficult to find suitable exposed populations in which to conduct hypothesis-driven gene–environment interaction studies that explore a putative link between genetic susceptibility and cancer risk from specific pesticides. Because of their lipid solubility and long biological half-life, there has been substantial interest in a putative link between organochlorine compounds (OCs), including PCBs and several older pesticides (notably DDT, dieldrin, aldrin, endrin, and hepatochlor), and breast cancer. In general, there has been little evidence of a significant association between OCs and breast cancer (Calle et al., 2002). However, when studies have considered common polymorphisms in certain drug metabolizing enzymes, in particular CYP1A1, some weak associations have been reported, suggesting that women with variant forms of CYP1A1 could be at slightly increased risk for breast cancer when PCB exposures are considered (Laden et al., 2002; Li et al., 2005; Zhang et al., 2004). Dozens, if not hundreds, of population-based and case–control studies examining associations between breast cancer risk and common genetic polymorphisms in xenobiotic metabolizing enzymes have been completed, but relatively few of these have explicitly considered pesticide exposures. One study found a significant association between a common genetic polymorphism in CYP1B1 and breast cancer risk in women living near hazardous waste incinerators for more than 10 years, or for women exposed during their life to agricultural products used in farming (Saintot et al., 2004), although no specific pesticides or exposure measures were considered. Thus, although there remains considerable interest in the hypothesis that genetic factors may predispose individuals to increased cancer risk from pesticide exposure – especially if exposures occur in development or early childhood (Van Larebeke et al., 2005) – there are very

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few studies with adequate exposure metrics and statistical power to demonstrate whether specific genetic polymorphisms, singly or combined, represent significant risk factors for specific types of cancer.

17.2.2  Pesticides, Genes, and Parkinson’s Disease Parkinson’s disease (PD) is the most common neurodegenerative disease of the elderly. This disease is characterized by a progressive loss of dopaminergic neurons in the substantia nigra pars compacta, resulting in characteristic motor symptoms, often accompanied by dementia. Numerous studies have reported both environmental and genetic contributors to the etiology of PD. The specific reasons why dopaminergic neurons seem to die with age at a more rapid rate in some people and not others is not well understood, although mitochondrial dysfunction coupled with characteristic protein aggregation (Lewy body formation) seem to be important. Genetics clearly plays a role in the etiology of PD, although highly penetrant mutations contribute to less than 10% of all PD cases, with most cases considered to be “sporadic” with no known etiology. Causative mutations in PD have been associated with loss of function variants for -synuclein (encoding SNCA; PARK1 and PARK4 loci on 4q21) (Migliore and Coppede, 2009). Other mutations in genes associated with ubiquitin processing of proteins (UCH-1: PARK5; Parkin: PARK2) and mitochondrial function (PINK-1: PARK6; DJ-1: PARK7; LRRK2: PARK8; ATP13A2: PARK9; OMI/ HTRA2: PARK13) (Migliore and Coppede, 2009) have also been linked to PD. Genetic studies in mono- and dizygotic twins support a modest genetic role in the etiology of sporadic (nonfamilial) PD. Indeed, numerous less penetrant “susceptibility genes” have been identified and perhaps function in concert with environmental factors. Among these are genes that encode proteins involved in dopamine transport (DAT, DRD2, COMT, MAO-B), xenobiotic metabolism (CYP2D6, GSTs, NAT2), and oxidative stress (NOX, SOD2). Given that genetics cannot alone explain most cases of PD, numerous environmental factors, including pesticides, have been examined in a large number of case–control and, more recently, several large cohort epidemiology studies. Among the various risk factors associated with PD, smoking is perhaps the most robust finding, although the effect is protective rather than causative (Elbaz and Tranchant, 2007). The reason why smoking appears consistently to be associated inversely with PD is unclear. Some studies have suggested that it may be due to the chronic stimulatory effect of nicotine, or nicotine-induced inhibition of MAOB, while others have attributed these observations to uncontrolled confounding (bias) secondary to differential rates

Chapter | 17  Genetic Polymorphism and Susceptibility to Pesticides

of smoking cessation in PD cases, or differences in smoking-related mortality. However, studies that have carefully addressed these concerns still find an inverse association between smoking and PD (Elbaz and Tranchant, 2007). Among the various environmental factors that have been implicated as causative factors in PD through epidemiology studies, agricultural practices, and specifically pesticides, are among the most commonly identified and have been reviewed extensively (Brown et al., 2006; Drechsel and Patel, 2008; Elbaz et al., 2007; Elbaz and Moisan, 2008; Elbaz and Tranchant, 2007; Hatcher et al., 2008; Le Couteur et al., 1999; Migliore and Coppede, 2009). Dick et al. (2007a,b) conducted a case–control study of 767 PD patients and 1989 matched controls that included detailed assessment of lifetime occupational/environmental exposures to a variety of substances, including solvents, pesticides, and metals. The only positive odds ratios (ORs) were for pesticide exposures [OR for high vs. nonexposed, 1.41; 95% confidence interval (CI), 1.06–1.88]. Another smaller population-based case–control study in Minnesota found that exposure to pesticides was associated with PD in men but not women (OR, 2.4; CI, 1.1–5.4; P  0.04). The association remained significant after adjustment for education or smoking. As with the studies by Dick et al., analyses for six other categories of industrial and household chemicals did not find any other significant associations. Although occupational and population-based studies have not definitely identified specific pesticides as etiological agents in the development of PD, there has been considerable interest focused on paraquat, maneb, and rotenone. Paraquat is a bipyridal quaternary ammonium compound that has found extensive use as a postemergent herbicide worldwide. It is best known toxicologically for causing serious, often fatal, pulmonary damage following extensive exposure (usually, but not always, suicide attempts). Paraquat readily undergoes redox cycling and accumulates specifically in lung tissue, apparently because of the presence of polyamine transporter systems in type II lung cells (Dinis-Oliveira et al., 2008). However, paraquat shares a remarkable structural resemblance to MPP  , the activated metabolite of 1-methyl-4-phenylpyridinium (MPTP). MPP is selectively toxic to dopaminergic neurons and is responsible for the development of a rapid-onset parkinsonism in young adults exposed to MPTP following use of adulterated methamphetamine (Langston et al., 1999). Because both MPP and paraquat cause substantial cytotoxicity in a variety of cellular and animal models via induction of oxygen free radicals, there is compelling “biological plausibility” for paraquat as a potential etiological factor for PD (Dinis-Oliveira et al., 2006, 2008). Indeed, numerous animal models of PD utilize paraquat to generate selective loss of dopaminergic neurons (Carvey et al., 2006; Cory-Slechta et al., 2005; Thiruchelvam et al., 2002;

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Thrash et al., 2007). However, it should be noted that there are distinct mechanistic differences in how paraquat, MPP  , and other neurotoxic compounds cause selective dopaminergic cell loss (Ramachandiran et al., 2007; Richardson et al., 2005). Despite the fact that several epidemiological studies have implicated pesticides, including herbicides, as risk factors for PD, and there is compelling mechanistic evidence that paraquat has the potential to cause selective toxicity to dopaminergic neurons, there is relatively little compelling epidemiological data demonstrating an association between paraquat exposure and PD (Allam et al., 2005; Brown et al., 2006; Dick et al., 2007b; Elbaz et al., 2007; Elbaz and Moisan, 2008; Hatcher et al., 2008). However, “absence of evidence” is not “evidence of absence,” so the putative link between paraquat and PD cannot be completely dismissed. Indeed, a few studies that included specific information on paraquat exposures have found positive associations. One of the earliest suggested associations between paraquat and PD was a small case��– control study in British Columbia, where a statistically significant association between PD development and paraquat contact (P  0.01) was reported (Hertzman et al., 1990). In another case–control study in Taiwan, Liou et al. (1997) found that the history of living in a rural environment, farming, use of herbicides/pesticides, and use of paraquat were associated with an increased PD risk in a dose– response relationship; the biological gradient between PD and previous uses of herbicides/pesticides and paraquat remained significant even after adjustment for multiple risk factors through conditional logistic regression. Further, the PD risk was greater among subjects who had used paraquat and other herbicides/pesticides than those who had used herbicides/pesticides other than paraquat. Another population-based case–control study found that a history of field crop farming, grain farming, herbicide use, or insecticide use resulted in a significantly increased crude estimate of the PD risk (Semchuk et al., 1992). In the multivariate analysis, which controlled for potential confounding or interaction between the exposure variables, previous occupational herbicide use was consistently the only significant predictor of PD risk. Another study reported an OR of 1.67 (CI, 0.22–12.76) for occupational exposure to paraquat, but it was based on only four cases (Firestone et al., 2005). The Agricultural Health Study examined both incident and prevalent cases of PD for many pesticides, including paraquat, and found an OR of 1.8 (95% CI, 1.0, 3.8) for paraquat and prevalent PD cases but an OR of 1.0 (95% CI, 0.5, 1.9) for incident cases of PD (Kamel et al., 2007). A GIS-based system that integrates California Pesticide Use reports and land-use maps was used to aid in pesticide exposure ascertainment among PD cases and matched controls (Costello et al., 2009). Potential residential exposures to paraquat, maneb, or the combination of the two was

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assessed, based on residence and GIS-based pesticide use information. When the groups as a whole were compared, there was no significant association between pesticide exposure and PD for either paraquat alone (OR 1.01; 95% CI, 0.71, 1.43) or maneb alone (OR 3.04; 95% CI, 0.30, 30.9), although there were only three cases and one control for maneb alone. However, when combined exposures were considered (those cases and controls with potential exposures to both pesticides) a significant association with PD was seen (OR, 1.75; 95% CI, 1.13, 2.73). The apparent effect was substantially greater (OR, 5.07; 95% CI, 1.75, 14.71) for those 61 years old at the time of diagnosis. Virtually all epidemiological studies examining possible associations between PD and individual pesticides suffer from limited size and thus weak statistical power, and very few have adequate measures of actual exposure. Thus, convincing epidemiologic data have failed to identify a specific and significant association between paraquat and PD, although such data are extremely difficult to obtain, given the relatively few people exposed repeatedly and the relative rarity of the disease in the population. These and other challenges in understanding the associations, or lack thereof, between specific pesticides and PD have been thoroughly described by Hatcher et al. (2008). There are few obvious genes that might be proposed to increase susceptibility to paraquat neurotoxicity. This fact, when combined with the somewhat limited use of paraquat (at least in the United States) and the difficulties in quantitatively estimating exposures, makes gene–environment interaction studies on the putative association between paraquat and PD exceedingly difficult to conduct. In addition to (and in combination with) paraquat, maneb has also been proposed as a potential risk factor for PD. The combination of paraquat and maneb has been used experimentally to induce PD-like symptoms in mice and is particularly effective when maneb is administered during fetal development, followed by challenge with paraquat in adulthood (Barlow et al., 2007). As noted previously, there is some epidemiologic evidence supporting the potential interaction between maneb and paraquat in the development of PD, as residential exposure to both pesticides appeared to have a stronger association with PD than either pesticide alone (Costello et al., 2009). However, as with paraquat alone, there are no obvious candidate “environmental susceptibility” genes that might interact with maneb exposures, and thus no studies have directly addressed this question. Rotenone is yet another pesticide that has been investigated for its possible contributions to PD susceptibility. The primary mechanism of action of rotenone is inhibition of mitochondrial function through high affinity binding and inhibition of complex I (NADH-dehydrogenase), which is also a target for MPP . However, unlike MPTP, rotenone readily diffuses across biological membranes and thus does not require membrane transporters for

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uptake into dopaminergic neurons. Nevertheless, rotenone is somewhat selectively toxic to dopaminergic neurons. Rotenone-induced dopaminergic lesions have cytoplasmic inclusions containing -synuclein, just as seen in human PD. Rotenone may also promote activation of microglia and subsequent neuroinflammation, which might contribute to the PD-like lesions that occur following rotenone injections (Gao et al., 2003). Although rotenone is widely used experimentally to induce PD-like pathology in rats (Ceravolo et al., 2006; Hoglinger et al., 2006; Miller et al., 2009; Schmidt and Alam, 2006), there is little evidence to suggest that its use as a pesticide contributes in any way to the etiology of sporadic PD. Rotenone is very poorly absorbed in the GI tract, has a relatively short environmental half-life, and its limited use in aquatic systems does not afford much opportunity for widespread human exposure (Hatcher et al., 2008). Thus, although there are few compelling mechanistically based “candidate genes” for which to study potential gene– environment interactions between specific pesticides and PD, many studies have utilized the candidate gene approach in attempts to identify putative environmental susceptibility genes for PD, without regard to specific pesticides or putative mechanisms. For example, associations between PD and SNPs in genes encoding CYP2D6; GSTs M1, P1 and T1; DAT, Mn-dependent SOD, and other genes have been completed and are discussed in detail in Section 17.4.

17.3  Importance of environmental exposure assessment in g  e studies of pesticides Perhaps the biggest challenge in the study of gene– environment interactions for pesticides is reliable, quantitative exposure assessment information. Mechanistic studies are of great value in identifying putative susceptibility genes. To demonstrate such a potentially significant genetic effect requires accurate knowledge of the actual exposed dose. Seldom is such information available. For the vast majority of population-based studies, exposures are often crudely estimated, usually based on proxy measures such as proximity to agricultural land, pesticide use records, or simply responses to questionnaires about pesticide use. At best, these provide qualitative estimates that allow categorical classifications such as “exposed” vs. “nonexposed.” Actual exposures within the “exposed group” may range by several orders of magnitude, making it difficult to imagine how a modest change in activity of a single enzyme might be detected against the noise of multiple different metabolic pathways and the uncertainty in actual dose. For some polymorphisms, such as those for the GSTM1 and GSTT1 homozygous null genotypes, the phenotypic difference (e.g., catalytic activity toward a given substrate) is large since activity is completely lacking in homozygous

Chapter | 17  Genetic Polymorphism and Susceptibility to Pesticides

null individuals. However, even these extreme phenotype examples do not abrogate the need for good exposure assessment since the presence or absence of a functional allele is still basically modifying the effectiveness of the exposed dose. Thus, the development of better measures of exposure is of paramount importance to the study of gene– pesticide interactions (Fenske, 2005; Infante-Rivard and Weichenthal, 2007; Metayer and Buffler, 2008). Some biomarkers of pesticide exposure, such as blood or plasma cholinesterase, are “effects related,” and are very useful clinically, but basically integrate the effects of the exposed dose (e.g., mg of systemically available, biologically active pesticide) with the sum of all genetic factors that modify the biological activity of the absorbed dose (e.g., rates of activation and detoxification by various biotransformation enzymes). Therefore, even effects-related biomarkers of pesticide exposure do not necessarily provide the type of exposure data that are useful in ascertaining whether a particular polymorphism in a pesticide disposition gene contributes to varying susceptibility. Exposure-related biomarkers, such as the direct measurement of the parent pesticide and/or metabolites of the pesticide in blood or urine, ostensibly provide the most useful information in terms of determining if a genetic polymorphism contributes to increased or decreased susceptibility to a given pesticide. But even here, there are significant challenges. For example, the quantitative measurement of trichlorpyridinol (TCPy) in the urine has been used as a quantitative biomarker of chlorpyrifos exposure for years, based on the assumption that all of the TCPy found in urine was derived from exposure to chlorpyrifos. But recent studies have demonstrated that a significant fraction – perhaps as much as 95% – of TCPy found in urine may actually be from exposure to TCPy itself on foodstuffs rather than from exposure to chlorpyrifos (Eaton et al., 2008). Direct measurement of the pesticide in blood is certainly the best marker of actual exposure. For example, a study of chlorpyrifos exposure from indoor use of chlorpyrifos was conducted in households in New York City by directly measuring the levels of chlorpyrifos in fetal and maternal blood (Whyatt et al., 2004, 2005, 2007). These studies used a remarkably sensitive and quantitative mass spectrometry-based technique (Barr et al., 2002) to identify trace levels of chlorpyrifos in both mothers and babies living in homes where chlorpyrifos had been used indoors. However, exposure estimates were based on only single samples taken in the hospital at the time of delivery, or shortly after, and thus it is uncertain how reflective the levels were of exposures that occurred weeks and months earlier, during the period of embryonic and fetal development. This is important in light of the fact that the plasma half-life of chlorpyrifos is measured in hours, rather than days, at least following acute, relatively high dose exposures (Nolan et al., 1984; Timchalk et al., 2002).

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Although quantitative exposure assessment techniques for pesticides remain a major challenge for studies attempting to identify gene–environment interactions with pesticides, new technologies for creating small, sensitive, and cost-efficient personal monitors for low-level chemical exposures may help to address the challenges of quantitative exposure assessment for pesticides (Barry et al., 2009). Unraveling the genetic component of variability in response to pesticide exposures will undoubtedly require widescale applications of such new technologies in the future. Finally, a major problem present in any type of epidemiology study attempting to find associations between pesticide exposures, genetic polymorphism, and diseases such as PD or cancers is the insidious onset and chronic nature of the disease. Thus, even with adequate exposure assessment tools, major challenges remain in that the exposures of interest and consequence to the development of a chronic disease may well have occurred decades earlier.

17.4  Specific genes and polymorphisms relevant to putative gene–pesticide interactions 17.4.1  Cytochromes P450 CYPs are membrane-bound heme-thiolate enzymes that are well-known for their ability to metabolize xenobiotics en route to their detoxification and removal from the body. These enzymes may also inadvertently function to metabolically activate compounds (i.e., procarcinogens). CYPs, which account for 70–80% of all phase I xenobiotic-metabolizing enzymes, may vary widely in their expression and activity between any two individuals. This variability may be attributed to both genetic and environmental factors (Wolff and Strecker, 1992). Pharmacogenetics studies have provided knowledge regarding CYPs that is relevant not only to drug metabolism and corresponding toxicity but also to susceptibility to pesticides. In addition to effectively activating a variety of drugs, several CYPs have proven to be proficient at metabolizing pesticide substrates (Table 17.1). A suitable first example is CYP2C9, which has been highly studied with regard to its role in the metabolism of warfarin, a rodenticide that, for many years, has been used therapeutically in humans as an anticoagulant. Polymorphisms in CYP2C9 and several other genes have proven to be very influential in determining appropriate doses of warfarin (Gage et al., 2008). Polymorphisms in CYP2D6 have been examined in the context of both pesticides and PD and have revealed interesting but ultimately equivocal results (Elbaz et al., 2004; Singh et al., 2008). CYPs 3A4 and 3A5 have been implicated as efficient metabolizers of organophosphorous pesticides, but the effect of individual polymorphisms in

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Table 17.1  Proteins and Enzymes Potentially Related to Pesticide Toxicity Protein/enzyme of interest

Pesticide substrate(s)

Reference(s)

CYP2C9

Warfarin

Khan et al. (2003) Bloch et al. (2002) Adcock et al. (2004) Malhi et al. (2004) Li et al. (2006) Wadelius and Pirmohamed (2007) Limdi et al. (2008)

Methoxychlor

Hu et al. (2004)

VKORC1

Warfarin

Malhi et al. (2004) Li et al. (2006) Wadelius and Pirmohamed (2007)

CYP2D6

Chlorpyrifos

Sams et al. (2000) Mutch et al. (2006)

Parathion

Sams et al. (2000) Mutch et al. (2003) Mutch et al. (2006)

Diazinon

Sams et al. (2000) Mutch et al. (2006)

Chlorpyrifos

Sams et al. (2000) Dai et al. (2001) Buratti et al. (2003) Mutch et al. (2006)

Parathion

Buratti et al. (2003) Mutch et al. (2003) Mutch et al. (2006)

Diazinon

Buratti et al. (2003) Mutch et al. (2006)

Aziphos-methyl

Buratti et al. (2003)

Malathion

Buratti et al. (2005)

Chlorpyrifos

Buratti et al. (2003) Mutch et al. (2006)

Parathion

Buratti et al. (2003) Mutch et al. (2003) Mutch et al. (2006)

Diazinon

Buratti et al. (2003) Mutch et al. (2006)

Malathion

Buratti et al. (2005)

Chlorpyrifos

Buratti et al. (2003) Mutch et al. (2006)

Parathion

Butch et al. (2003)

CYP3A4/5

CYP1A2

CYP2B6

Mutch et al. (2006)

CYP2C8

Diazinon

Buratti et al. (2003) Mutch et al. (2006)

Aziphos-methyl

Buratti et al. (2003)

Malathion

Buratti et al. (2005)

Parathion

Mutch et al. (2003) Mutch et al. (2006)

Chapter | 17  Genetic Polymorphism and Susceptibility to Pesticides

531

Table 17.1  (Continued) Protein/enzyme of interest

Pesticide substrate(s)

Reference(s)

CYP2C8

Chlorpyrifos

Mutch et al. (2006)

Diazinon

Mutch et al. (2006)

Parathion

Mutch et al. (2006)

Chlorpyrifos

Mutch et al. (2006)

Diazinon

Mutch et al. (2006)

CYP2E1

Dibromochloropropane

Au et al. (1999)

PON1

Parathion

Adkins et al. (1993) Humbert et al. (1993) Taylor et al. (2000) Costa et al. (2003)

Diazinon

Davies et al. (1996) Cherry et al. (2002) Mackness et al. (2003)

Chlorpyrifos

Cole et al. (2003) Cole et al. (2005) Furlong et al. (2005)

Dibromochloropropane

Au et al. (1999)

Carboxylesterase (mouse and human)

Pyrethroids

Ghiasuddin and Soderlund (1984) Leng et al. (1999) Stok et al. (2004) Anand et al. (2006a, b) Crow et al. (2007) Yang et al. (2009)

GSTM1

1,3-Dichloropropene

Vos et al. (1991)

DAT

Heptachlor

Purkerson-Parker et al. (2001)

Permethrin

Gillette and Bloomquist (2003) Elwan et al. (2006)

Avermectin pesticides

Umbenhauer et al. (1997) Lankas et al. (1998) Mealey et al. (2002) Nelson et al. (2003)

CYP2C19

PGP

their corresponding genes is not yet clear. CYPs 1A1, 1A2, 1B1, 2B6, 2E1, 2C8, and 2C19 have also been studied to varying degrees for their potential contributions to pesticide toxicity and pesticide-related diseases; some of these relationships have been observed to be minor and/or highly specific (Table 17.1). During the past two decades, the relationship between CYP2C9 and the rodenticide warfarin has been well established. Initially marketed as a rat and mouse poison, warfarin has been widely used in the United States as an anticoagulant since the 1950s. Pharmacogenetic studies have revealed that polymorphisms in CYP2C9 as well as vitamin K epoxide reductase complex 1 (VKORC1) significantly affect individual susceptibility to warfarin. CYP2C9*1

is classified as the wild-type genotype. CYP2C9*2 and CYP2C9*3 encode proteins with 12 and 5% of the activity of the wild-type CYP2C9 protein, respectively (Khan et al., 2003) (Table 17.2). With regard to VKORC1, two major haplotypes (A and B) have been identified that affect warfarin sensitivity; individuals carrying the A haplotype have increased sensitivity to the drug. Both CYP2C9 and VKORC1 variants contribute to variability in effective dose, although the contribution of VKORC1 is more substantial. A case study found that the presence of two CYP2C9 variants (CYP2C9*2 and CYP2C9*3) enhanced the susceptibility of a 90-year-old patient to warfarin (Bloch et al., 2002). A subsequent study indicated similar sensitivity in two elderly women who were homozygous CYP2C9*3

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532

Table 17.2  Genetic Polymorphisms Potentially Involved in Human Susceptibility to Pesticide Exposures and/or Disease Gene(s)

Variant polymorphism(s)

Pesticide exposure

Disease/health effect

Reference

CYP2C9

CYP2C9*2, *3

Warfarin

Bleeding Bleeding Bleeding Bleeding

Khan et al. (2003) Bloch et al. (2002) Adcock et al., 2004 Malhi et al. (2004)

CYP2C9*2, 3

Methoxychlor

Hu et al. (2007)

VKORC1*497T   G, *698C  T, *1173C  T, *1542G  C, *2255T  C, *3730G  A; CYP2C9*2, *3

Warfarin

Li et al. (2006)

CYP2C9*2, *3, *5, *6, *10, *11; VKORC1 1173 C/T

Warfarin

Limdi et al. (2008)

VKORC1–1639/3673 G  A; CYP2C9*2, *3

Warfarin

Gage et al. (2008)

CYP2D6A, B, L, (PM)

N/A

PD

Bandmann et al. (1997)

CYP2D6*193G  A

N/A

PD

Chan et al. (1998)

Multiple

N/A

PD

Joost et al. (1999)

CYP2D6 (PM)*3, *4

N/A

Neurodegenerative disorders

Nicholl et al. (1999)

CYP2D6 29B

General

PD with dementia

Hubble et al. (1998)

CYP2D6 (PM) *3, *4, *5

General

PD

Deng et al. (2004)

CYP2D6*4

N/A

PD

Santt et al. (2004)

CYP2D6*4

N/A

PD

Singh et al. (2008)

VKORC1 and CYP2C9

CYP2D6

CYP1A1

CYP2E1

PON1

CYP1A1m1, m2, m4

General

Childhood leukemia

Infante-Rivard et al. (1999)

CYP1A1*2 A, 2B

N/A

Neurodegenerative disorders

Nicholl et al. (1999)

CYP1A1*2

N/A

Childhood ALL

Canalle et al. (2004)

CYP1A1m1

N/A

Non-Hodgkin’s lymphoma

Kerridge et al. (2002)

CYP1A1*2C

General

da Silva et al. (2008)

CYP2E1m

General

Au et al. (1999)

CYP2E1*5B

General

da Silva et al. (2008)

CYP2E1*3

N/A

Childhood ALL

Canalle et al. (2004)

CYP2E1 G1259C

N/A

Childhood ALL

Krajinovic et al. (2002)

CYP2E1*5B, *6

N/A

PD

Singh et al. (2008)

Q192R

Parathion

Humbert et al. (1993)

Q192R

Diazinon

Davies et al. (1996)

Q192R, L55M

Parathion

Mackness et al. (1997)

PON1 coding and promoter polymorphisms

N/A

Brophy et al. (2001a,b)

Q192R, L55M, -108C/T

Organophosphates

Costa et al. (2003)

Q192R

Organophosphates

Furlong et al. (2006)

PON1 coding and promoter polymorphisms

General

Berkowitz et al. (2004)

-108C/T and Q192R

Organophosphates

Childhood brain tumors

Searles Nielsen et al. (2005)

Chapter | 17  Genetic Polymorphism and Susceptibility to Pesticides

533

Table 17.2  (Continued) Gene(s)

Variant polymorphism(s)

Pesticide exposure

Disease/health effect

Reference

Q192R

General

Male reproductive outcomes

Padungtod et al. (1999)

Q192R

Organophosphates

Male reproductive effects

Perez-Herrera et al. (2008)

Q192R, L55M

Diazinon in sheep dip

Ill health

Cherry et al. (2002)

Q192R, L55M

Diazinon in sheep dip

Ill health

Mackness et al. (2003)

Q192R

General

Various

Lee et al. (2003)

-909G/C, -108C/T, L55M, Q192R

General

Hernandez et al. (2003)

Q192R

General

Hernandez et al. (2004)

Q192R

Organophosphates

Lee et al. (2007)

Q192R

General

Lopez et al. (2007)

Q192R, L55M

N/A

PD

Clarimon et al. (2004)

L55M

General

PD

Fong et al. (2005)

L55M

N/A

PD

Akhmedova et al. (2001)

L55M

N/A

PD

Carmine et al. (2002)

Multiple

General

PD

Benmoyal-Segal et al. (2005)

Q192R

N/A

Non-Hodgkin’s lymphoma

Kerridge et al. (2002)

Q192R, L55M

Organophosphates

-108C/T, L55M, Q192R

N/A

Q192R, L55M

General

Q192R, L55M

N/A

Brain lesions

Hadjigeorgiou et al. (2007)

PON1 coding and promoter polymorphisms

General

ALS

Morahan et al. (2007)

Q192R, L55M

N/A

AD

Leduc and Poirier (2008)

Q192R, L55M

Organophosphates

Zhou et al. (2007)

Q192R, L55M, -108C/T

Organophosphates

Sirivarasai et al. (2007)

-909G/C, -162A/G, -108C/T, L55M, Q192R in PON1

N/A

Chen et al. (2003)

PON1 and PON2

Q192R in PON1; S311C in PON2

General

PD

Taylor et al. (2000)

PON2 and PON3

Multiple

N/A

Sporadic ALS

Saeed et al. (2006)

BChE

BChE-K

Organophosphates

Various

Zhou et al. (2007)

GSTM1

GSTM1 null

1,3-Dichloropropene

Vos et al. (1991)

GSTM1 null

General

Gregio D’Arce and Colus 2000

GSTM1 null

N/A

Neurodegenerative disorders

Nicholl et al. (1999)

GSTM1 null

N/A

PD

Santt et al. (2004)

GSTM1 null

N/A

PD

Perez-Pastene et al. (2007)

GSTM1 null

N/A

Childhood ALL

Krajinovic et al. (2002)

Sozmen et al. (2002) Autism

D’Amelio et al. (2005) Browne et al. (2006)

(Continued)

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534

TABLE 17.2 (Continued) Gene(s)

Variant polymorphism(s)

Pesticide exposure

Disease/health effect

Reference

GSTP1

GSTP1*B

N/A

PD

Vilar et al. (2007)

GSTP1 I105V

General

Wong et al. (2008)

GSTO1 coding polymorphisms

Arsenic

Marnell et al. (2003)

Multiple

Arsenic

Ahsan et al. (2007)

GSTM1 and GSTP1

GSTM1 null and GSTP1 I105V

Arsenic

Marcos et al. (2006)

GSTM1 and GSTT1

GSTM1 null and GSTT1 null

General

Scarpato et al. (1996) (1997)

GSTM1 null and GSTT1 null

General

Au et al. (1999)

GSTM1 null and GSTT1 null

General

Falck et al. (1999)

GSTM1 null and GSTT1 null

General

Hernandez et al. (2005)

GSTM1 null and GSTT1 null

General

Kirsch-Volders et al. (2006)

GSTM1 null and GSTT1 null

N/A

PD

Bandmann et al. (1997)

GSTM1 null and GSTT1 null

General

Childhood leukemia

Infante-Rivard et al. (1999)

GSTM1 null and GSTT1 null

N/A

Non-Hodgkin’s lymphoma

Kerridge et al. (2002)

GSTM1 null and GSTT1 null

N/A

PD

Ahmadi et al. (2000)

GSTP1 and GSTT1

Multiple

N/A

PD

Singh et al. (2008)

GSTM1, GSTP1, and GSTT1

Multiple

Arsenic

Skin lesions

McCarty et al. (2007)

Multiple

General

Liu et al. (2006)

Multiple

General

da Silva et al. (2008)

Multiple

N/A

Childhood ALL

Canalle et al. (2004)

GSTM1, GSTP1, GSTT1, and GSTZ1

Multiple

General

PD

Menegon et al. (1998)

GSTM1, GSTM3, GSTP1, and GSTT1

Multiple

General

PD

Dick et al. (2007a,b)

DAT1 (SCL6A3)

DAT1 10/9

N/A

Neurodegenerative disorders

Nicholl et al. (1999)

DAT1 1215 A/G

N/A

PD

Lin et al. (2002)

DAT1 1215 A/G

N/A

PD

Singh et al. (2008)

Multiple

General

PD

Kelada et al. (2006)

MDR1

MDR1 C3435T

General

PD

Drozdzik et al. (2003)

NQO1

NQO1*2

General

PD

Dick et al. (2007a,b)

NQO1*2, *3

N/A

Childhood ALL

Krajinovic et al. (2002)

NQO1 C609T

N/A

Pediatric neoplasms

Kracht et al. (2004)

NQO1 C609T

N/A

Childhood leukemia

Smith et al. (2005)

GSTO1

N/A, not assessed; general, mixed or no specific type/class; PD, Parkinson’s disease; AD, Alzheimer’s disease; ALS, amyotrophic lateral sclerosis; ALL, acute lymphoblastic leukemia.

and heterozygous CYP2C9*2, respectively (Khan et al., 2003). Additional studies have called for increased clinician attention to CYP2C9 genotype when determining therapeutic doses of warfarin for patients (Adcock et al., 2004;

Gage et al., 2008; Limdi et al., 2008; Malhi et al., 2004). Several studies have concurrently addressed the additive contribution of CYP2C9 and VKORC1 polymorphisms to warfarin susceptibility. Three VKORC1 polymorphisms

Chapter | 17  Genetic Polymorphism and Susceptibility to Pesticides

were associated with warfarin dose, a finding that did not hold for the two major CYP2C9 polymorphisms (Li et al., 2006). However, a variant CYP2C9 genotype conferred significantly increased risk for major hemorrhage while VKORC1 1173 C/T did not (Limdi et al., 2008). Despite these discrepancies, CYP2C9 and VKORC1 are both thought to be important determinants of warfarin sensitivity (Wadelius and Pirmohamed, 2007). Less is known about the role of these polymorphisms/genes/proteins in the metabolism of other pesticides. Hu et al. (2004) suggest a role of CYP2C9 in methyoxchlor metabolism, but additional studies are needed. Like CYP2C9, much of the characterization of CYP2D6 expression and activity has been derived from laboratory and clinical investigations of a CYP2D6 specific substrate, debrisoquine. CYP2D6 is highly polymorphic, possessing over 80 characterized allelic variants (Daly, 2004) (Table 17.2). Genetic polymorphisms in CYP2D6 result in enzymes which have no, little, normal, or increased activity toward debrisoquine. From these investigations, genotype– phenotype correlations have been established. Individuals may now be categorized as poor, extensive, or ultra-rapid metabolizers based on CYP2D6 genotype. These categories have proven useful in studies of drug toxicity, and they have provided the rationale for studying CYP2D6 polymorphisms in the context of pesticide exposure. In vitro studies of CYP2D6 and other CYPs using human liver microsomes have shown that CYP2D6 effectively metabolizes chlorpyrifos, diazinon, and parathion (Mutch et al., 2003; Mutch and Williams, 2006; Sams et al., 2000). Prior to and following these conclusions, a number of groups have examined CYP2D6 polymorphisms and their relationship to PD susceptibility, given the hypothesized link between pesticide exposure and PD. In the absence of pesticide exposure data, epidemiologic studies of Caucasian and Chinese populations indicated that there is no significant association between CYP2D6 polymorphisms and PD (Bandmann et al., 1997; Chan et al., 1998). Joost and colleagues chose to investigate CYP2D6 more closely and screened the CYP2D6 gene for additional polymorphisms. None of these polymorphisms differed in frequency between PD cases and controls. Similarly, they found no association between the polymorphisms and family history of PD, age of onset, or environmental exposures (Joost et al., 1999). A study of the contribution of polymorphisms in CYP2D6, CYP1A2, N-acetyltransferase 2, dopamine active transporter, and glutathione S-transferase M1 to neurodegenerative diseases (sporadic PD, familial PD, ALS, multiple system atrophy, progressive supranuclear palsy, and Alzheimer’s disease) indicated that none of the suspected polymorphisms were associated with these diseases (Nicholl et al., 1999). Despite these null findings, some groups have found evidence of a link between CYP2D6 polymorphisms, pesticide exposure, and PD risk. A study by Hubble et al. demonstrated that subjects with at least one copy of the “poor metabolizer” 2D6 allele who

535

had been exposed to pesticides had an 83% predicted probability of PD plus dementia. These conclusions were based on a relatively small sample (43 PD  D and 51 PD – D patients) (Hubble et al., 1998). Two subsequent studies also found evidence of potentiation between CYP2D6 genotype and pesticide exposure, an interaction which was positively associated with PD risk (Deng et al., 2004; Elbaz et al., 2004). These findings suggest that in the presence of pesticides, individuals who are poor metabolizers are at greatest risk for toxicity. This is somewhat surprising given that the metabolites of OP pesticides that are generated by CYP2D6 metabolism are often more toxic than their parent compounds, at least in terms of cholinesterase inhibition, which infers that ultra-rapid or extensive metabolizers would be at the greatest risk of pesticide toxicity. (However, there is little evidence to suggest any association between chronic inhibition of acetylcholinesterase and development of PD, so the putative role of CYP2D6 polymorphisms contributing to differences in susceptibility to PD in OP-exposed populations is speculative, at best.) This discrepancy as well as the discordance of the existing studies indicate that the relationship between CYP2D6 polymorphisms, pesticide toxicity, and PD risk is still unclear. More recent studies have suggested a link between members of the CYP3A family and pesticides, particularly those of the organophosphorous (OP) family. CYPs 3A4, which is expressed in the liver and gastrointestinal tract, and 3A5, whose expression is both hepatic and extrahepatic, together contribute to the metabolism of more than 50% of all pharmaceuticals. CYP3A7 expression exceeds that of all other CYP3A enzymes at fetal stages, but its expression decreases over time while CYP3A4 correspondingly increases. CYP3A enzymes have broad substrate specificity. For this reason, they may also be capable of processing numerous xenobiotics (i.e., pesticides) or endogenous factors, in addition to metabolizing a host of drugs. Expression of CYP3A4/5 (and others, including CYP2C8) is highly variable, varying by as much as 150fold (Mutch and Williams, 2004). Genetic variation is suspected to account for 70–90% of the interindividual differences in constitutive expression and activity of CYP3A enzymes (Ozdemir et al., 2000). Several polymorphisms in the coding region of CYP3A4 have been reported, including 18 that result in frameshift or nonsynonymous mutations; however, these changes are considered rare and have somewhat limited effects on catalytic activity (IngelmanSundberg et al., 2001). The most common CYP3A4 polymorphism is located in the 5 region of the gene and has shown association with increased transcription of the gene in vitro, results that have not been demonstrated in the human population (Ball et al., 1999). Unlike CYP3A4, variants in CYP3A5 appear to be more strongly associated with phenotype. The most “important” CYP3A5 polymorphism is an intronic SNP (CYP3A5*3), A6986G, which results in the production of a truncated, nonfunctional version of the CYP3A5 protein. This polymorphism,

536

CYP3A5*3, is actually the common allele; other minor alleles include CYP3A5*1, CYP3A5*5, CYP3A5*6, and CYP3A5*7. Individuals with at least one CYP3A5*1 allele express high levels of CYP3A5 (Table 17.1). It is estimated that CYP3A5 is present and active in 10–30% of livers, a frequency which may vary by ethnicity. A number of studies have assessed the roles of CYP enzymes in the metabolism of OPs and have identified CYPs 3A4 and3A5 as major players (Buratti et al., 2003, 2005; Dai et al., 2001; Mutch et al., 2003; Mutch and Williams, 2006; Sams et al., 2000). Mutch and Williams (2006) examined the ability of human liver microsomes (HLMs) and recombinant CYP enzymes to metabolize diazinon, chlorpyrifos, and parathion. As expected, wide variation in activity toward these substrates was observed among the individual microsome samples. This may be due to a variety of factors, including genetics, nutrition, and medications. Data from these samples collectively suggested, however, that CYPs 3A4 and 3A5 are extensively involved in both the activation and the detoxification of diazinon to diazoxon and pyrimidinol, respectively. CYPs 2C8 and 2C19 preferentially generated the detoxification metabolite. With regard to chlorpyrifos, production of the chlorpyrifos detoxification product, TCPy, correlated with the activity of CYPs 3A4/5, 2C8, 2C19, and 1A2 in the HLM samples. With the exception of CYP2C19, the same CYPs were also involved in the production of paraoxon from parathion. Ultimately, using a combination of HLM and recombinant CYP data, the authors concluded that CYPs 3A4/5, 2C8, 1A2, 2C19, and 2D6 all play roles in the metabolism of the three studied OPs, and that both exposure level/duration and the presence of polymorphisms are likely to be important in determining individual susceptibility to these pesticides (Mutch and Williams, 2006). Additionally, CYP2B6 is also particularly adept at activating chlorpyrifos to chlorpyrifos oxon, although its role in the metabolism of other OPs is minor or negligible. These data represent an overarching principle in xenobiotic metabolism: Often, multiple enzymes are capable of metabolizing multiple substrates. For this reason, strongly linking a single CYP polymorphism to a particular pesticide remains a challenging endeavor. Additional CYPs have been examined to determine their involvement in pesticide toxicity and pesticide-associated diseases. CYP1A1 also metabolizes hexachlorobenzene (Hahn et al., 1988), warfarin (Zhang et al., 1995), and parathion (Mutch et al., 1999). CYP1A1 polymorphisms have been assessed for their effects on susceptibility to diseases such as childhood acute lymphoblastic leukemia, PD, and non-Hodgkin’s lymphoma, yielding results that were either speculative or null (Canalle et al., 2004; InfanteRivard et al., 1999; Nicholl et al., 1999; Schroeder, 2005). The majority of these studies also included pesticide exposure as a variable. CYP1A2 is also capable of metabolizing warfarin and further plays a role in OP metabolism. Thus, variability in CYP1A2 activity and/or expression

Hayes’ Handbook of Pesticide Toxicology

may influence individual susceptibility to these pesticides (Buratti et al., 2003, 2005; Mutch et al., 2003; Mutch and Williams, 2006). CYPs 1A2, 2B6, and 2C19 have primary roles in OP metabolism at low levels of substrate, while CYP3A4 becomes more active at high levels, illustrating the complexity of mammalian biotransformation systems. Lastly, studies of pesticide-exposed farm workers in Costa Rica and Brazil have reported associations between CYP2E1 genotype (specifically, a polymorphism in the upstream regulatory region that results in increased CYP2E1 expression) and DNA damage, as measured by cytogenic and genotoxic assays (Au et al., 1999; da Silva et al., 2008). (The Brazilian epidemiologic study included additional polymorphisms in GSTs; these will be covered in Section 17.4.4). Variability in the expression and/or activity of CYP enzymes is attributed to a number of sources, one of which is genetic variation. Interindividual differences in the ability to process pesticides may be partially explained by this source of variability, although the contributions of genetics are controversial. In many cases, CYP genotype does not correlate well with CYP phenotype, a phenomenon which is likely due to the fact that these enzymes are induced to varying degrees by external factors such as drugs, lifestyle factors (e.g., smoking, alcohol intake), and components of the diet. Additionally, compensatory mechanisms are often at play in xenobiotic metabolism, processes that may be beneficial or detrimental, depending on the enzyme and the substrate. Nonetheless, the previously described studies demonstrate that (1) CYPs play a role in pesticide toxicity and (2) polymorphisms in the genes encoding these enzymes have the potential to affect susceptibility to pesticides and related diseases.

17.4.2 Paraoxonase PON, an enzyme that protects low-density lipoprotein against high-density lipoprotein peroxidation, also hydrolyzes paraoxon, the active form of parathion. Genetic variability in PON function was first reported in 1983 (Eckerson et al., 1983). Ten years later, a PON1 polymorphism (Gln192Arg) was identified that partially distinguishes individuals with high and low activity toward paraoxon (Adkins et al., 1993; Humbert et al., 1993). The Arg192 (R192) isoform of PON1 rapidly hydrolyzes paraoxon, yielding protection to those who possess this polymorphism (Table 17.2). This generalization does not hold for other toxic compounds, including diazinon, soman, and sarin, which are more efficiently detoxified by individuals with the Gln192 (Q192) isoform of PON1 (Davies et al., 1996). In addition to Q192R, a polymorphism at amino acid site 55 (Leu55Met) in PON1 has also been investigated (Table 17.2; some studies use position 55, others use 54). In a study of 279 healthy human subjects, individuals who expressed the Q192 isoform of PON1 and were homozygous for M

Chapter | 17  Genetic Polymorphism and Susceptibility to Pesticides

at the 55 position were proposed to be the most susceptible to OP poisoning (Mackness et al., 1997). MM homozygotes proved to have the lowest PON1 activity, a trait that was independent of the 192 genotype. Variability in PON1 expression is an important contributor to PON1 function, and thus the PON1 genotype alone is not fully indicative of phenotype (Brophy et al., 2000). Phenotyping assays are thus generally more reliable to identify “PON1 status” of an individual. Several new approaches using non-OP substrates have been developed for the rapid assessment of PON1 status in individuals (Huen et al., 2009; Richter et al., 2009). In addition to the widely studied Q192R and L55M PON1 polymorphisms, polymorphisms in the PON1 promoter as well as in PON2 have also been investigated for their contributions to PON1 expression and activity levels, pesticide metabolism, and susceptibility to pesticide-related conditions (Taylor et al., 2000). Three promoter polymorphisms (909, 162, and 108) were identified that increased the expression level of PON1 (Brophy et al., 2001a) (Table 17.2). The −108C/T polymorphism accounted for approximately 23% of the variability in PON1 protein levels (Brophy et al., 2001b). Two additional noncoding polymorphisms have been reported, but the –108 polymorphism remains the most relevant to PON1 levels (Costa et al., 2003). Polymorphisms in the noncoding regions of PON1 and PON2 may contribute to pesticide susceptibility, but they have received relatively less attention than those in the coding regions of these enzymes. PON1 levels vary not only within age strata but also between age groups. Expression levels of PON1 are relatively low during development and plateau between 6 and 15 months of age in humans (Cole et al., 2003; Furlong et al., 2000). These studies were the first to suggest that neonates may have an increased susceptibility to pesticides due to their PON1 status (Cole et al., 2003; Furlong et al., 2000). In a genetic epidemiologic study of Caucasian, Caribbean Hispanic, and African-American neonates and their mothers, three PON1 promoter and two coding polymorphisms were investigated for their effects on expression and activity levels. The activity levels of the neonates were significantly lower than those of the mothers (Chen et al., 2003). Another study of Latina mothers and their newborns showed that, on average, PON1 levels of the children were fourfold lower than those of the mothers (Furlong et al., 2006). In a prospective cohort study of mothers and infants in New York City, pesticide exposure, which was based on questionnaire data and pesticide metabolite levels, and PON1 status were examined with respect to their effects on weight, length, head circumference, and gestational age. Maternal chlorpyrifos levels in combination with low maternal PON1 activity levels were associated with reduced head circumference (Berkowitz et al., 2004), although the relevance of the chlorpyrifos exposures (vs. the PON1 polymorphism itself) to the observed effects is not clear (Eaton et al., 2008). Yet another study found no

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significant link between -108C/T or Q192R and risk for childhood brain tumors, but these conclusions were based on a relatively small sample size (66 cases and 236 controls) (Searles Nielsen et al., 2005). Studies conducted to date definitively show that PON1 levels differ significantly within individuals and between children and adults, a property that is likely very relevant in the context of susceptibility to pesticides. Since the characterization of functional polymorphisms in PON1 and the identification of vulnerable populations, numerous population-based studies have been conducted among pesticide-exposed individuals. These studies have investigated the relationship between PON1 genotype, pesticide exposure, and various health effects, including male reproductive outcomes. A study of Chinese pesticide factory workers indicated that unexposed Q192 homozygotes and exposed R192 homo/heterozygotes had significantly lower sperm counts and fewer morphologically normal sperm compared with unexposed R192 homozygotes. These conclusions were based on a small sample size and should be interpreted with caution, but they suggest that PON1 genotype modifies the effect of OP exposure on male fertility (Padungtod et al., 1999). This was supported by another study in Mexican farmers (Perez-Herrera et al., 2008). Among a sample of farmers exposed to sheep dip, which contains diazinon as its active ingredient, selfreported ill individuals were more likely to be heterozygous or homozygous R allele at position 192 of PON1 (OR, 1.93; 95% CI, 1.24, 3.01) and homozygous LL at position 55 (OR, 1.70; 95% CI, 1.07, 2.68) compared with “healthy” controls (Cherry et al., 2002; Mackness et al., 2003). A similar association was observed by Lee and colleagues in a study of pesticide-exposed South African farmers (Lee et al., 2003). Both coding and noncoding polymorphisms were investigated by Hernandez et al. in a group of 102 greenhouse workers suspected to be exposed to a variety of pesticides. These individuals were compared with nonsprayer controls. Although no significant differences were observed between specific genotypes and exposure to pesticides, the authors found that protective clothing positively influenced PON1 activity (Hernandez et al., 2003). Follow-up studies indicated that esterase activity was decreased in pesticide applicators compared to nonapplicators (Hernandez et al., 2004, 2005), an effect that was observed in other long-term pesticide exposure studies (Zhou et al., 2007). The relationship between PON1 genotype, antioxidant enzyme levels, and oxidative stress is another avenue under investigation (Lee et al., 2007; Lopez et al., 2007). PON genotypes have been studied not only in the context of pesticide exposures but also in the context of disease. Pesticide exposure may cause acute, subchronic, and chronic toxicity, but the majority of published studies have been devoted to neurodegenerative diseases and cancers. Although the association between pesticide exposure

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and risk of PD has been reported in numerous studies (see Section 17.2.2), the associations between PON genotypes and PD have been less consistent. Several studies in different populations (Australian, Chinese, Finnish, and Taiwanese) have observed no association between PON1/2 genotypes and PD risk (Clarimon et al., 2004; Fong et al., 2005; Taylor et al., 2000; Wang and Liu, 2000). Other studies, using cases and controls from Russian and Swedish populations, respectively, have presented evidence of an association between PON1 polymorphism L55M and PD; specifically, individuals who carry the M allele were reported to be at significantly increased risk of PD compared to those with the L allele (Akhmedova et al., 2001; Carmine et al., 2002). The interaction between acetylcholinesterase and paraoxonase polymorphisms may be important to the risk of pesticide-induced PD (BenmoyalSegal and Soreq, 2006; Benmoyal-Segal et al., 2005). Other health outcomes have been examined with regard to pesticide exposure and PON polymorphisms, including non-Hodgkin’s lymphoma, autism, brain lesions/tumors, sporadic amyotrophic lateral sclerosis, Alzheimer’s disease, and general effects on reproduction, development, and functioning (Akgur et al., 2003; Browne et al., 2006; Cummings and Kavlock, 2004; D’Amelio et al., 2005; Hadjigeorgiou et al., 2007; Kerridge et al., 2002; Leduc and Poirier, 2008; Morahan et al., 2007; Saeed et al., 2006; Schroeder, 2005; Searles Nielsen et al., 2005; Sozmen et al., 2002). Many of these reports suggest that PON genotypes modify the association between pesticide exposure and health outcomes, but small sample sizes and lack of generalizability, including specificity of pesticides and lack of exposure data, continue to be an issue in disease–PON1 association studies. Since it is often ethically and financially infeasible to conduct studies of controlled exposures and polymorphisms in human subjects, mouse models provide a plausible alternative. Furlong and colleagues observed that mice lacking PON1 are highly sensitive to chlorpyrifos oxon and diazoxon but not to paraoxon (Costa et al., 2003; Furlong et al., 1998; Li et al., 2000). Injecting these knockout mice or wild-type mice with PON1 protein (particularly the R192 isoform in the case of chlorpyrifos) protected against the effects of chlorpyrifos oxon and diazoxon. Additional studies in wild-type mice have demonstrated that adult levels of PON1 are achieved by approximately 3 weeks of age; prior to this, PON1 levels are significantly lower, an observation that also holds true for human infants (Furlong et al., 2000). Cole and colleagues generated transgenic mice that lacked endogenous PON1 and expressed human PON1 (either Q192 or R192). These mice exhibited similar PON1 expression levels with respect to time as wild-type mice, which verified the conservation of regulatory sequences between human and mouse PON1 (Cole et al., 2003). In subsequent studies, they illustrated the functionality of the PON1 coding polymorphism by exposing the

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mice to chlorpyrifos and chlorpyrifos oxon in vivo. Mice that expressed the human PON1 Q192 isoform were more sensitive to both chlorpyrifos and chlorpyrifos oxon than mice expressing the human R192 isoform (Cole et al., 2005; Furlong et al., 2005). Given the sequence-related and functional similarities of PON1 between humans and mice, these findings were revealing. They provided additional insight into the role of PON1 in determining individual susceptibility to pesticides, particularly chlorpyrifos. The preceding review summarizes the wealth of research that has been done with regard to PON1 genotypes, pesticide exposures, and disease associations. Both mouse and human studies have convincingly demonstrated the role of PON1 polymorphisms in susceptibility to pesticide toxicity. The case still needs to be made for diseases that are putatively associated with chronic pesticide exposure. This continues to be an area of interest, one which will be explored further with both mouse models and human populations.

17.4.3  Additional Carboxylic Esterases As indicated by the volume of literature described previously, PON1 is the most highly studied polymorphic esterase with regard to pesticide toxicity. Additional carboxylic esterases, such as carboxylesterases and butyrylcholinesterases, may also play a role in pesticide susceptibility. Early studies documented the ability of mouse carboxylesterases to hydrolyze pyrethroid insecticides, a detoxification system that is similar to that of PON1 (Ghiasuddin and Soderlund, 1984; Stok et al., 2004). Individual variability in carboxylesterase expression and activity was later proposed by Leng and colleagues as a factor in susceptibility to pyrethroid pesticides (Leng et al., 1999). Studies of deltamethrin (DLM), a type II pyrethroid, using rat liver microsomes and plasma, illustrated the participation of both carboxylesterases and CYPs in its detoxification (Anand et al., 2006a). Immature rats experienced increased neurotoxicity when exposed to DLM compared to adult rats (Anand et al., 2006b). Like PON1, carboxylesterase levels appear to be lower at earlier stages of development, a property which may confer increased risk of health effects following pyrethroid exposure (Yang et al., 2009). Humans and rodents differ in their serum content of carboxylesterases and thus may exhibit differential susceptibility to pyrethroids (Crow et al., 2007; Godin et al., 2006). The major human carboxylesterases (hCE1 and hCE2) possess different affinities for different pyrethroid substrates and are expressed at varying levels in relevant organs. The highest activities of hCE1 and hCE2 are in the liver. hCE2 is also the predominant carboxylesterase in human intestine, and it is highly active toward transpermethrin but not DLM or bioresmethrin. These specificities coupled with the inherent variability in carboxylesterase expression between individuals may serve to explain differential

Chapter | 17  Genetic Polymorphism and Susceptibility to Pesticides

susceptibility to pyrethroid toxicity (Crow et al., 2007). Yang and colleagues examined the expression levels of hCE1 and hCE2 in human liver samples from adults, children, and fetuses and showed that carboxylesterase expression is positively correlated with age. They also observed a huge variability within age groups, suggesting the potential involvement of polymorphisms or other modifying factors (Yang et al., 2009). As referenced in a review by Ross and Crow (2007), human hepatic samples are variable in their hydrolytic activities, a property that does not appear to correlate with hCE expression. This discrepancy might be due to the presence of polymorphisms that affect enzyme activity. Given the affinity of hCEs for pyrethroids, especially those of the type II class, investigating the genetic variation in human carboxylesterases is a necessary next step in predicting individual susceptibility to pyrethroids and other pesticides. Variants in butyrylcholinesterase (BChE) have been well-studied in the context of diseases such as Alzheimer’s disease and type 2 diabetes, but they are also highly relevant to pesticide toxicity. Over 65 variants have been detected in BChE (Table 17.2), but the BChE K polymorphism has received the most attention (Mikami et al., 2008). The K variant is relatively common compared to the other known variants, and it causes a reduction of ~30% in the activity of the resulting protein (Bartels et al., 1992). In a study of 75 workers exposed to OPs (and 100 nonexposed controls), both the BChE K genotype and the major PON1 variant genotypes exhibited more severe symptoms than those with wild-type genotypes among those who were exposed to OPs (Zhou et al., 2007). Regardless of genotype, both BChE and carboxylesterase appear to be effectively inhibited by OPs. A notable decrease in BChE activity was observed in individuals who had been poisoned with OP pesticides, compared to unexposed controls (Sirivarasai et al., 2007; Sozmen et al., 2002). The same effect was observed in the study by Zhou and colleagues. These data collectively suggest a role for additional carboxylic esterases in susceptibility to pesticides.

17.4.4  Glutathione S-Transferase GSTs participate in the phase II metabolism of xenobiotics by catalyzing the conjugation of glutathione (GSH), an antioxidant peptide, to reactive metabolites, which often are derived from phase I reactions. GSH conjugation may occur in the absence of GSTs and/or may involve the parent compound rather than the metabolite, but enzymatically mediated GSH reactions are more common. GST reactions are usually classified as detoxification processes, although on rare occasions, GSH conjugates may be toxic or reactive (Hayes et al., 2005). A number of classes of GSTs have been characterized in human tissues; these include alpha (GSTA), kappa (GSTK), mu (GSTM), pi (GSTP), omega (GSTO), sigma

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(GSTS), theta (GSTT) and zeta (GSTZ). For many classes, there are multiple different genes, with individual genes within a class designated by a number, given in order of discovery (e.g. GSTM1, GSTM2, GSTM3). All but GSTK are localized to the cytosol, although several cytosolic GSTs have also been found in mitochondria (Raza et al., 2002). GSTs are equipped to metabolize both xenobiotics and endogenous compounds; their substrates include drugs, environmental pollutants, and products of oxidative stress. Several GSTs have been examined with respect to pesticide susceptibility, including GSTs M1, O1, P1, T1, and Z1 (Table 17.1). All of the human cytosolic GSTs are polymorphic, and many of the polymorphisms result in increased or decreased GST protein expression (Table 17.2). In the case of both GSTM1 and GSTT1, particular alleles (GSTM1*0 and GSTT1*0) are absent due to a gene deletion. The frequencies of homozygous null polymorphisms (GSTM*0/*0 and GSTT1*0/*0) have been estimated to be approximately 50 and 20%, respectively, in Caucasians, although the frequencies can vary substantially by ethnic group (Eaton, 2000). These two polymorphisms have been given the most attention in disease association studies, but a number of other polymorphic GST variants have also been identified (Table 17.2). Three additional polymorphisms have been documented for GSTM1, including two polymorphisms that change an amino acid (GSTM1*A and GSTM1*B) and another, GSTM1*1  2, which causes overexpression of GSTM1 (Hayes et al., 2005). Four GSTP1 polymorphisms (GSTP1*A, B, C, and D) change amino acids in their corresponding proteins. The same is true for GSTZ1, which also possesses four allelic variants (GSTZ1*A, B, C, and D). Six polymorphisms have been documented for GSTO1 and O2. Over 20 studies have been conducted to evaluate the relationship between GSTs and pesticide toxicity. One of the first of these studies examined the ability of GSTM1 positive (those individuals possessing one or two functional alleles) and null individuals to metabolize the soil fumigant 1,3-dichloropropene (DCP) (Vos et al., 1991). Studies by Scarpato and colleagues then focused on the role of GSTM1 and GSTT1 deletion polymorphisms in DNA damage [chromosomal aberrations (CA)] among pesticide-exposed floriculturists and greenhouse workers. Pesticide exposure alone was not associated with increases in DNA damage, but, among smokers, individuals who were homozygous null for GSTM1 had statistically higher CA frequencies than those who had at least one copy of GSTM1 (Scarpato et al., 1996, 1997). GSTM1–GSTT1 double null individuals also had increased CA frequencies compared to GSTT1 null individuals, but the sample size was quite low (n  5). These initial studies made a case for the role of GST polymorphisms in detoxification of the components of cigarette smoke but not for pesticides. Subsequent studies presented conflicting results regarding the contribution of GST polymorphisms to DNA damage

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following pesticide exposure. Au and colleagues observed correlations between GSTM1 and T1 deletion polymorphisms and cytogenetic effects of dibromochloropropane; the “unfavorable” alleles for CYP2E1 and PON1 also showed a significant relationship with these effects (Au et al., 1999). However, another study reported that genotoxicity, as evidenced by increases in micronuclei (MN), was higher among GSTM1 positive pesticide-exposed workers compared to GSTM1 null individuals, which was an unexpected finding (Falck et al., 1999). No statistically significant association between CA frequencies and GSTM1 genotype was found among Brazilian agricultural workers exposed to pesticides, although, again, the sample size was limiting (Gregio D’Arce and Colus, 2000). Due to these disparities, early reviews cautioned the premature use of GST polymorphisms in predicting disease risk associated with exposure to pesticides (Eaton, 2000; Sram, 1998). More recent studies have employed larger sample sizes and focused on the collective effects of GSTM1 and T1 polymorphisms as well as variants in GSTP1 and O1. Several other studies have examined three GST polymorphisms, alone or in combination, in pesticide-exposed workers. Hernandez and colleagues (2005) showed that the null genotypes for GSTM1 and T1 were predictive of pesticide-related symptoms in a study of pesticide applicators and controls. Association between these null genotypes and DNA damage in another cohort study of pesticideexposed agricultural workers was also reported (da Silva et al., 2008). Additional studies have demonstrated that GSTP1 variant genotypes are associated with increased risk of DNA damage among pesticide-exposed fruit growers (da Silva et al., 2008; Hernandez et al., 2005; Liu et al., 2006; Wong et al., 2008). Kirsch-Volders and colleagues reported somewhat contradictory findings, indicating that among individuals nonoccupationally and occupationally exposed to genotoxic substances, those lacking GSTT1 exhibited lower MN frequencies, evidence of a protective effect (Kirsch-Volders et al., 2006). In general, however, the body of evidence suggests that variant/null GST genotypes confer increased susceptibility to pesticide-induced toxicity. Given the proposed linkage between GST genotypes and pesticide-related health effects, a number of studies have investigated the link between these genotypes and disease, including PD, acute lymphoblastic leukemia (ALL), and non-Hodgkin’s lymphoma (NHL), but were typically conducted without regard to specific pesticide exposures or any exposure assessment. Although early studies did not find strong associations between GST genotype and PD risk (Bandmann et al., 1997; Menegon et al., 1998; Nicholl et al., 1999; Paolini et al., 1999), more recent findings suggest that wild-type versions of GSTM1, T1, and P1 are generally associated with lower PD risk and/or later age of onset (Ahmadi et al., 2000; Dick et al., 2007a; Perez-Pastene et al., 2007; Santt et al., 2004;

Hayes’ Handbook of Pesticide Toxicology

Singh et al., 2008; Vilar et al., 2007). With regard to ALL and NHL, in general, GST polymorphisms only appeared to be associated with disease in the presence of other “atrisk” alleles in genes such as PON1, NQO1, and CYP2E1 (Canalle et al., 2004; Infante-Rivard et al., 1999; Kerridge et al., 2002; Krajinovic et al., 2002). Polymorphisms in genes encoding GSTs have effectively been linked with disease, as indicated by the studies mentioned previously. Deficiencies in these proteins have already been shown to decrease the protection of individuals from the acute and chronic health outcomes that may be caused by exposure to pesticides. Given their role in the general detoxification of environmental toxicants, GSTs may contribute to defense against the harmful effects of pesticides. However, there is a general paucity of studies with both adequate exposure assessment for specific pesticides and adequate statistical power to demonstrate statistically significant associations between specific GST polymorphisms, specific pesticides, and specific diseases. The lack of such studies is ���������������������������������������� not surprising�������������������������� , however, given the challenges in identifying large populations with substantial and documented pesticide exposures and prevalence of disease, which must then be further stratified by genotypes.

17.4.5  Additional Putative Pesticide Susceptibility Genes In addition to the esterases (i.e., PON1) and the well-known phase I and II biotransformation enzymes described previously, there are other candidate genes that may influence an individual’s susceptibility to pesticides. The dopamine transporter (DAT) should be among the first to be considered. As its name suggests, DAT is responsible for transporting dopamine from the synaptic cleft into dopaminergic neurons. This action effectively terminates the dopamine signal. The gene encoding DAT is polymorphic, having at least 63 variants (Greenwood et al., 2006). Many of these genetic variants have been examined for putative relationships to a variety of neurological diseases and conditions, including bipolar disorder, attention deficit hyperactivity disorder, and PD. Studies of DAT polymorphisms in PD have yielded equivocal results (Higuchi et al., 1995; Kim et al., 2000; Lin et al., 2002, 2003; Nicholl et al., 1999; Nishimura et al., 2002; Singh et al., 2008), but many of these did not include consideration of pesticide exposures. In a study of PD patients and controls with and without pesticide exposure, Kelada et al. (2006) showed that “risk” alleles in SLC6A3, the gene that encodes DAT, are associated with PD and that pesticide exposure may further modify this risk. The rationale for studying the role of DAT in pesticide susceptibility stems from research in the early 1980s of 1-methyl-4-phenyl-1,2,3,6-tetrahydropyridine (MPTP). MPTP exposure was found to cause a Parkinson’s-like syndrome; this condition was later connected to its toxic

Chapter | 17  Genetic Polymorphism and Susceptibility to Pesticides

metabolite MPP (Langston and Ballard, 1983; Langston et al., 1983, 1984). Human DAT is very effective at transporting MPP into dopaminergic neurons, where it may then exert its toxic effects (Lee et al., 1996; Richardson et al., 2005). As discussed in Section 17.2.2, paraquat, a herbicide, is structurally very similar to MPP  , and thus it has been proposed that paraquat may function similarly to MPP  . Several groups have suggested that paraquat toxicity involves DAT (Ossowska et al., 2005; Yang and Tiffany-Castiglioni, 2005), but studies by Richardson et al. (2005) indicate that MPP  , paraquat, and rotenone all function distinctly with regard to DAT. Regardless, paraquat appears to have some interaction with DAT and thus the investigation of the role of DAT polymorphisms in paraquat metabolism and susceptibility is warranted. In addition to paraquat, the insecticides permethrin and heptachlor have also been shown to affect DAT expression and binding. Heptachlor increased dopamine transport and DAT expression in exposed C57BL/6 mice (Miller et al., 1999). Related experiments in Sprague–Dawley rats showed increases in DAT binding at all stages of development, even at low levels of heptachlor (Purkerson-Parker et al., 2001), and these changes appeared to persist throughout the life cycle. Like heptachlor, permethrin exposures in mice have caused a persistent upregulation of DAT as well as -synuclein, which is a major component of Lewy bodies, protein agglomerates found in brains of PD patients (Gillette and Bloomquist, 2003). In addition to upregulating DAT, permethrin exposure can induce apoptosis (Elwan et al., 2006). Together, these studies indicate that certain pesticide exposures could potentially make individuals increasingly susceptible to subsequent pesticide exposures (or endogenous dopamine-related toxicity) due to the persistent upregulation of DAT expression and/ or binding. DAT polymorphisms are likely to contribute to this susceptibility and are worthy of consideration. Like DAT, MDR1, which encodes P-glycoprotein (PGP), may affect pesticide susceptibility. In this case, however, PGP is involved in transporting toxic substances out of cells rather than taking them in. Studies have shown that a particular strain of mice, CF-1, is highly sensitive to avermectin pesticides due to a lack of PGP (Lankas et al., 1998; Umbenhauer et al., 1997). This susceptibility has also been observed in dogs, specifically collies (Mealey et al., 2002; Nelson et al., 2003). Drozdik and colleagues applied this knowledge to the human population, investigating the MDR1 C3435T polymorphism in relation to pesticide exposure and PD. Although no statistical association was observed between the polymorphism and PD, there was such an association between the variant and PD patients who had been exposed to pesticides. Unfortunately, as is often the case with epidemiologic studies, the specific pesticide or pesticides to which the study subjects were exposed were not noted, only information regarding method and duration of exposure. Exposed PD patients had

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an increased frequency of the polymorphism compared to unexposed patients, and the presence of one or both copies of the variant allele conferred an increased risk of disease (Drozdzik et al., 2003). A review of the potential functionality of polymorphisms in human PGP (ABCB1) that included specific discussion on the genetic susceptibility of humans to avermectins concluded that “it is likely that humans carrying at least one functional copy of ABCB1 will not be more susceptible to avermectin toxicity at clinically relevant doses or at the low exposure levels resulting from pesticide use” and further that homozygotes for a loss of function haplotypes for ABCB1 would be “very rare in human populations” (Macdonald and Gledhill, 2007). Given its ability to detoxify a variety of substrates, NAD(P)H:quinone oxidoreductase 1 (NQO1) has been investigated with regard to pesticide exposures and related diseases, including PD. The major NQO1 variant that has received attention is the C609T polymorphism, which decreases NQO1 activity (Table 17.2). Although one study did not find an interaction between NQO1 polymorphisms, pesticide exposure, and PD (Dick et al., 2007a), another found that pesticide exposure was significantly associated with PD in the Taiwanese population and that a combination of variant genotypes in NQO1 and manganesecontaining superoxide dismutase (MnSOD) was associated with increased risk for PD among pesticide-exposed subjects (Fong et al., 2007). NQO1 polymorphisms have also been examined in the context of leukemia and lymphoma, although the role of pesticide exposures in this scenario has yet to be explored (Kracht et al., 2004; Krajinovic et al., 2002; Smith et al., 2004, 2005).

Conclusions The Human Genome Project has generated an unprecedented opportunity, as well as the necessary tools and technologies, to explore how individual genetic variability can influence the chances of developing a pesticide-related disease. Although much of the past research on gene– environment interactions regarding pesticides and disease focused on one or a few specific genetic polymorphisms, it is becoming increasingly evident that the strength of association between any individual genetic variant and the disease of interest (i.e., penetrance) is likely to be low (e.g., ORs typically less than 1.2) and thus very difficult to identify. It is likely that most genetic susceptibility arises not through the effect of a single variant but, rather, an unwelcome combination of multiple variants, each contributing a small amount to risk. Technologies are now in hand to measure hundreds of thousands of specific genetic variants (usually as “single nucleotide polymorphisms”) in an individual, although it remains costly to apply such extensive analyses to the thousands of samples needed for a statistically robust study. New high-throughput sequencing

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approaches will make it possible within a few years to sequence the entire genome of an individual for less than $1000. Such extensive genetic knowledge of individuals in a population now allows for “genome-wide association studies” (GWAS) to study the relationship between genetic variability and specific diseases. GWAS are currently being used to identify new genetic markers of disease, although their use in gene–environment interaction studies remains problematic, largely because of inadequate exposure assessment. This is especially true when one is looking for associations between pesticide exposures, genetic susceptibility factors, and chronic diseases such as cancer, Parkinson’s or Alzheimer’s diseases, cardiovascular disease, etc., simply because the exposures of interest likely occurred in the distant past. Recent interest in the role of prenatal and early childhood exposures to nutritional and other environmental factors as a basis for adult diseases [so-called Barker hypothesis (Barker, 2007; de Boo and Harding, 2006; Sinclair et al., 2007)] makes exposure assessment a particularly challenging problem. Since most “mechanistically based” hypotheses about specific genetic polymorphisms as a basis for increased susceptibility to a particular pesticide are focused on genetic variants in metabolism/disposition genes, knowledge of the level of exposures over time is particularly important, since most such variants simply move an individual “up����������������������������������������������������� ” or ���������������������������������������������� “down” the dose–response curve. Failure to accurately quantify dose in individuals makes exposure misclassification a particularly serious problem. Thus, it seems paramount that future studies aimed at elucidating genetic risk factors for pesticide-related diseases consider 1. Multiple potential genetic variants, rather than one or a handful of candidate genes 2. Detailed consideration of pesticide exposures, including consideration of early life exposures, and specific pesticides Just these two criteria alone provide a huge obstacle to such studies – statistical power, and the challenge of false discovery rates that arise from multiple comparisons (Martin et al., 2007; Storey and Tibshirani, 2003). Thus, the new approaches and technologies ushered in by the human genome project require a paradigm shift in how future gene–environment interaction studies on pesticides are done. Such studies will almost certainly have to be highly collaborative and dispersed, e.g., multiple groups working together on different populations, with common protocols and careful bioinformatic and statistical approaches incorporated into the design of the studies. Perhaps the best example of such a future study is the National Children’s Study, which is a prospective, national cohort study that will examine the effects of environmental influences on the health and development of 100,000 children across the United States, following them from before birth until age 21 (http://www.nationalchildrensstudy.gov/Pages/default.aspx). A second example is the

Agricultural Health Study (AGH; http://aghealth.nci. nih.gov/), also funded by a consortium federal agencies (led by the National Institute of Occupational Health and safety). This study, which began in 1994, explores potential causes of cancer and other diseases among farmers and their families and among commercial pesticide applicators. The AGH is designed to identify occupational, lifestyle, and genetic factors that may affect the rate of diseases in farming populations (Alavanja et al., 1996). Finally, the Northern California Childhood Leukaemia Study represents another prospective cohort study that includes both early childhood exposure assessment and extensive analysis of genetic polymorphism (Metayer and Buffler, 2008). The challenge with this and most other localized cohorts will be the statistical limitations imposed by the somewhat limited sample size in nested disease-association studies, especially for somewhat rare diseases. However, it is likely that other cohort studies will adopt similar methods, providing the opportunity for future meta-analyses of multiple cohorts, with greatly improved statistical power.

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Chapter | 17  Genetic Polymorphism and Susceptibility to Pesticides

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Thrash, B., Uthayathas, S., Karuppagounder, S. S., Suppiramaniam, V., and Dhanasekaran, M. (2007). Paraquat and maneb induced neurotoxicity. Proc. West Pharmacol. Soc. 50, 31–42. Timchalk, C., Nolan, R. J., Mendrala, A. L., Dittenber, D. A., Brzak, K. A., and Mattsson, J. L. (2002). A physiologically based pharmacokinetic and pharmacodynamic (PBPK/PD) model for the organophosphate insecticide chlorpyrifos in rats and humans. Toxicol. Sci. 66, 34–53. Umbenhauer, D. R., Lankas, G. R., Pippert, T. R., Wise, L. D., Cartwright, M. E., Hall, S. J., and Beare, C. M. (1997). Identification of a P-glycoprotein-deficient subpopulation in the CF-1 mouse strain using a restriction fragment length polymorphism. Toxicol. Appl. Pharmacol. 146, 88–94. Van Larebeke, N. A., Birnbaum, L. S., Boogaerts, M. A., Bracke, M., Davis, D. L., Demarini, D. M., Hooper, K., Huff, J., Kleinjans, J. C., Legator, M. S., Schoeters, G., and Vahakangas, K. (2005). Unrecognized or potential risk factors for childhood cancer. Int. J. Occup. Environ. Health 11, 199–201. Vilar, R., Coelho, H., Rodrigues, E., Gama, M. J., Rivera, I., Taioli, E., and Lechner, M. C. (2007). Association of A313 G polymorphism (GSTP1*B) in the glutathione-S-transferase P1 gene with sporadic Parkinson’s disease. Eur. J. Neurol. 14, 156–161. Vos, R. M., van Welie, R. T., Peters, W. H., Evelo, C. T., Boogaards, J. J., Vermeulen, N. P., and van Bladeren, P. J. (1991). Genetic deficiency of human class mu glutathione S-transferase isoenzymes in relation to the urinary excretion of the mercapturic acids of Z- and E-1,3dichloropropene. Arch. Toxicol. 65, 95–99. Wadelius, M., and Pirmohamed, M. (2007). Pharmacogenetics of warfarin: current status and future challenges. Pharmacogen. J. 7, 99–111. Wang, J., and Liu, Z. (2000). No association between paraoxonase 1 (PON1) gene polymorphisms and susceptibility to Parkinson’s disease in a Chinese population. Mov. Disord. 15, 1265–1267. Whyatt, R. M., Rauh, V., Barr, D. B., Camann, D. E., Andrews, H. F., Garfinkel, R., Hoepner, L. A., Diaz, D., Dietrich, J., Reyes, A., Tang, D., Kinney, P. L., and Perera, F. P. (2004). Prenatal insecticide exposures and birth weight and length among an urban minority cohort. Environ. Health Perspect. 112, 1125–1132. Whyatt, R. M., Camann, D., Perera, F. P., Rauh, V. A., Tang, D., Kinney, P. L., Garfinkel, R., Andrews, H., Hoepner, L., and Barr, D. B. (2005). Biomarkers in assessing residential insecticide exposures during pregnancy and effects on fetal growth. Toxicol. Appl. Pharmacol. 206, 246–254. Whyatt, R. M., Garfinkel, R., Hoepner, L. A., Holmes, D., Borjas, M., Williams, M. K., Reyes, A., Rauh, V., Perera, F. P., and Camann, D. E. (2007). Within- and between-home variability in indoor-air insecticide levels during pregnancy among an inner-city cohort from New York City. Environ. Health Perspect. 115, 383–389. Wolff, T., and Strecker, M. (1992). Endogenous and exogenous factors modifying the activity of human liver cytochrome P-450 enzymes. Exp. Toxicol. Pathol. 44, 263–271. Wong, R. H., Chang, S. Y., Ho, S. W., Huang, P. L., Liu, Y. J., Chen, Y. C., Yeh, Y. H., and Lee, H. S. (2008). Polymorphisms in metabolic GSTP1 and DNA-repair XRCC1 genes with an increased risk of DNA damage in pesticide-exposed fruit growers. Mutat. Res. 654, 168–175. Yang, D., Pearce, R. E., Wang, X., Gaedigk, R., Wan, Y. J., and Yan, B. (2009). Human carboxylesterases HCE1 and HCE2: ontogenic expression, inter-individual variability and differential hydrolysis of oseltamivir, aspirin, deltamethrin and permethrin. Biochem. Pharmacol. 77, 238–247.

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Yang, W., and Tiffany-Castiglioni, E. (2005). The bipyridyl herbicide paraquat produces oxidative stress-mediated toxicity in human neuroblastoma SH-SY5Y cells: relevance to the dopaminergic pathogenesis. J. Toxicol. Environ. Health A 68, 1939–1961. Zhang, Y., Wise, J. P., Holford, T. R., Xie, H., Boyle, P., Zahm, S. H., Rusiecki, J., Zou, K., Zhang, B., Zhu, Y., Owens, P. H., and Zheng, T. (2004). Serum polychlorinated biphenyls, cytochrome P-450 1A1 polymorphisms, and risk of breast cancer in Connecticut women. Am. J. Epidemiol. 160, 1177–1183.

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Zhang, Z., Fasco, M. J., Huang, Z., Guengerich, F. P., and Kaminsky, L. S. (1995). Human cytochromes P4501A1 and P4501A2: R-warfarin metabolism as a probe. Drug Metab. Dispos. 23, 1339–1346. Zhou, Z. J., Zheng, J., Wu, Q. E., and Xie, F. (2007). Carboxylic esterase and its associations with long-term effects of organophosphorus pesticides. Biomed. Environ. Sci. 20, 284–290.

Chapter 18

Pesticides as Endocrine-Disrupting Chemicalsa Tammy E. Stoker and ��Robert ������� J. ����� Kavlock ��������� U.S. Environmental Protection Agency, Research Triangle Park, North Carolina

18.1  Introduction Hormones are chemical messengers secreted by endocrine glands that travel in the bloodstream or the fluid surrounding the cells to specific receptors within or on the surface of target cells. The binding of the hormone to the receptor initiates a specific response from the cell, by either altering the cell’s existing proteins or turning on genes that will build a new protein. To date, researchers have identified more than 50 hormones in humans and other vertebrates. Unfortunately, more than a few environmental substances are also known to bind unintentionally to hormone receptors and can imitate or block endogenous endocrine signaling. Yet other chemicals are known to modulate hormone synthesis and secretion or to alter metabolism and clearance. Collectively these materials are now known as “endocrine disruptors.” Hormones play a crucial role in guiding normal cell differentiation in early life-forms and so exposure to endocrine-disrupting substances during these critical developmental periods can cause effects that are not evident until later in life, such as effects on behavior and reproduction, and increased susceptibility to cancer or disease. The endocrine system consists of a number of central and peripheral organs (e.g., hypothalamus–pituitary, thyroid, parathyroid, adrenal, pancreas, ovaries, and testes) that synthesize, store, and release hormones (e.g., gonadotropins, thyroid hormone, parathyroid hormone, corticosterone, insulin, estrogen, progesterone, and testosterone) a 

Disclaimer: The research described in this article has been reviewed by the National Health and Environmental Effects Research Laboratory, U.S. Environmental Protection Agency and approved for publication. Approval does not signify that the contents necessarily reflect the views and policies of the Agency, nor does mention of trade names or commercial pro­ ducts constitute endorsement or recommendation for use.

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into the blood (Griffin and Ojeda, 1988; Hadley, 1996). These hormones, in turn, regulate the function of remote organs and tissues in the body to maintain homeostasis either by inducing (or suppressing) the synthesis of genes or by altering signal transduction pathways within the target cells. Most endocrine organs are linked with the hypothalamus and pituitary gland in classical negative feedback loops, which allow precise control of hormone levels in the blood. Frequently (especially during critical developmental periods) the levels of hormones are regulated within narrow limits by the feedback loops, but at other times (e.g., during the estrous cycle) large fluctuations in hormone levels are the intended result, as they interact with target tissues to either augment or repress the release of other hormones. Given the central role of the endocrine system in regulating homeostasis and controlling developmental processes, it is not surprising that interference with their normal action leads to alterations in either function or morphology. Indeed, pharmaceutical agents are often developed for such properties, be it the regulation of ovulation by birth control pills, the reduction of breast cancer risk by antiestrogens (e.g., tamoxifen) or aromatase inhibitors (e.g., anastrazole), or the reduction of prostate growth by anti­androgens (e.g., finasteride). Although the ability of pesticides to interact with endocrine function has been known for over 40 years, it was only in the 1990s that interest in this mode of action rose to a high level of prominence. For example, the estrogenic action of some DDT analogues was first reported in 1952 (Fisher et al., 1952), and a number of publications appeared in the 1970s demonstrating that the insecticides kepone (Gellert et al., 1978; Guzelian, 1982) and methoxychlor (Bulger et al., 1978) were estrogenic. With the reports of reproductive tract cancers and other disorders noted in the offspring of women who received diethylstilbestrol (DES, 551

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a powerful synthetic estrogen) in the 1950s to help prevent miscarriages (Herbst, 1972) and the follow-up work in animal models by others (Korach and McLachlan, 1985; McLachlan and Newbold, 1987), the phrase “environmental estrogens” became prevalent. In 1992, the World Wildlife Fund held a conference that examined a broad range of indicators of adverse health outcomes in humans and in wildlife (Colborn and Clements, 1992). Since then, many endocrine-disrupting chemicals (EDCs) have been shown to affect reproduction in wildlife populations. While some EDCs disappear quickly from the natural environment, others persist (based on physiochemical properties of the EDC, such as lipophilicity) and these have been the most studied. Aquatic animals are particularly affected, especially carnivores, because they are at the top of the “food chain” where high levels of persistent chemicals build up over time. Some examples of effects in wildlife include the following: l l l l l l

Reduction in the population of Baltic seals Eggshell thinning in birds of prey Decline in the alligator population in a polluted lake Reduction in frog populations Adverse effects on fish reproduction and development Development of male sex organs in female marine animals such as whelks and snails

For example, the effects in seals, birds, and alligators are most likely due to EDCs such as PCBs, dioxins, DDT/DDE, and other pesticides that contain chlorine. The effects on fish appear to be caused by estrogens in the water flowing from sewage treatment works into rivers. The effects on marine whelks and snails are presumed to be due to the use of TBT (tributyltin) in anti-fouling paints on boats and ships. A presumed pesticide spill in Lake Apopka, Florida, provides a well-publicized example of potential EDC effects on population decline in alligators. A variety of gonadal and developmental abnormalities were observed that have been attributed to high levels of various organochlorine contaminants that disrupt endocrine homeostasis. The effects on the alligators appear to be the result of exposure to chemicals resulting from a spill of the pesticide dicofol (or its metabolic breakdown products). These effects include reproductive failure leading to reductions in the number of neonate and juvenile offspring, developmental abnormalities of the reproductive tract and male phallus, and abnormal sex steroid levels (Guillette et al., 1994, 2007). Since 1995, there has been an explosion of workshops, symposia, publications, and committee efforts to define an “endocrine disruptor” as more than a compound that affects estrogen function and to identify the adverse effects mediated by exogenous chemicals on the endocrine system (Kavlock et al., 1996). The identification of wildlife populations experiencing adverse effects either on individuals

or on populations from areas contaminated with endocrinedisrupting chemicals combined with the observation of declines in human health indices such as sperm quality and cancers of the endocrine-regulated organs, such as the breast, testes, and prostate, have further raised concerns (Cooper and Kavlock, 1997; Kavlock and Ankley, 1996). The concern about potential adverse effects of endocrine disruptors culminated in the enactment of two laws by the U.S. Congress and included the requirement for screening chemicals for estrogenic and other endocrine activity (i.e., the Food Quality Protection Act of 1996 and the Safe Drinking Water Act of 1996). As a result of this legislation, the U.S. Environmental Protection Agency (U.S. EPA) established the Endocrine Disruptor Screening and Testing Advisory Committee (EDSTAC) and later the Endocrine Methods and Validation Subcommittee (EDMVAC) to assist in its implementation (see Section 18.2; http://www.epa. gov/scipoly/oscpendo/index.htm). The EDSTAC-recommended approach to evaluating potential endocrine-disrupting effects involving the estrogen, androgen, and thyroid hormone signaling pathways (U.S. EPA, 1998b) is arguably the largest new testing program to be proposed. In 2002, the World Health Organization’s International Programme on Chemical Safety held a meeting titled Global Assessment of the State-of-the-Science of Endocrine Disruptors (http://www.who.int/ipcs/publications/new_issues/endocrine_disruptors/en/). In its report, it concluded that “although it is clear that certain environmental chemicals can interfere with normal hormonal processes, there is weak evidence that human health has been adversely affected by exposure to endocrine-active chemicals. However, there is sufficient evidence to conclude that adverse endocrine-mediated effects have occurred in some wildlife species.” Citing the fact that studies to date examining EDC-induced effects in humans have yielded inconsistent and inconclusive results, the group wrote that although that explains their characterization of the evidence as weak, “[that] classification is not meant to downplay the potential effects of EDCs; rather, it highlights the need for more rigorous studies.” This “global assessment” further states that the only evidence showing that humans are susceptible to EDCs is currently provided by studies of high exposure levels. There is, in fact, clear evidence that intrauterine EDC exposures can alter human reproductive tract development and physiology. The most thoroughly characterized example is DES, the synthetic estrogen prescribed to millions of pregnant women in the United States and elsewhere from the 1940s to the 1970s to prevent miscarriage. The drug is known to have caused a rare form of vaginal cancer in thousands of daughters of women who took DES, as well as a variety of adverse reproductive tract effects in both the daughters and the sons of those women. In light of the complexity of the endocrine system and the multiple points at which it can be perturbed by exo­ genous agents, it is not surprising that endocrine disruption

Chapter | 18  Pesticides as Endocrine-Disrupting Chemicals

may be caused by a number of different chemical classes and structures, including a number of pesticides. One common underlying theme is that the reproductive system, particularly that of the developing organism, is especially vulnerable to the toxicity manifest by alterations in endocrine function; a number of pesticidal agents have been shown to exert their effects by this mode of action. Some pesticides, such as the insect growth regulators, are specifically developed for those abilities (although these do not appear to be endocrine disruptors in vertebrates), whereas for others the endocrine effects exhibited by target and nontarget organisms are clearly secondary to their primary mode of toxicity. In this chapter, four basic modes of action were used to classify endocrine-disrupting pesticides: (1) the ability to interact directly with steroid receptors; (2) the ability to modify steroid hormone metabolizing enzymes; (3) the ability to perturb hypothalamic–pituitary release of trophic hormones; and (4) the ability to alter thyroid hormones. The following text provides examples of the mechanisms of endocrine function which can be perturbed by chemical exposure and a few examples of pesticides which act via these mechanisms. It does not cover all aspects of their toxicity, as that is amply covered in other sections of this handbook. Impacts of altered endocrine function on development and reproduction, particularly in experimental animal models, will be emphasized. Studies using in vitro systems to detect modes of action will be mentioned only in conjunction with in vivo applications docu­menting that the mode of action is operable in an intact multicellular organism.

18.2  Steroid hormones 18.2.1  Estrogen Receptor Function 18.2.1.1  Methoxychlor Methoxychlor [1,1,1-trichloro-2,2-bis(4-methoxyphenyl) ethane; MXC] is a chlorinated hydrocarbon insecticide which was first used in the United States as a replacement for DDT [1,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane]. MXC exposure caused a number of estrogenic effects, such as vaginal opening and cornification (Cummings and Metcalf, 1994; Eroschenko, 1991), stimulation of uterine growth and hypertrophy (Swartz et al., 1994; Walters et al., 1993), increases in uterine peroxidase (Walters et al., 1993), and ornithine decarboxylase activities in rats (Bulger et al., 1978). While MXC is weakly estrogenic, the predominant bioactivity occurs upon its liver-mediated metabolism to 2,2bis(p-hydroxyphenyl)-1,1,1-trichloroethane (HPTE) (Gaido et al., 1999; Kupfer and Bulger, 1987; Nelson et al., 1978). Experimental studies have shown that HPTE binds to the estrogen receptor (ER) with higher affinity than methoxychlor. It has been reported that HPTE functions as an ER

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alpha (ER) agonist and ER beta (ER) antagonist (Gaido et al., 1999) and induces AR-dependent changes in gene expression following in vivo exposure in mice (Waters et al., 2001). The bis-hydroxy metabolites were even more potent in vitro. Methoxychlor binds with equal affinity to ER and ER with a binding affinity relative to 17-estradiol of 0.01 (Krisfalusi et al., 1998; Laws et al., 2000a). In addition to the estrogenic potential of methoxychlor and its metabolite, Maness et al. (1998) demonstrated that both MXC and HPTE in the range of 108 to 104 M reduced the activity of co-administered dihydrotestosterone but showed no agonist activity. The metabolite was approximately 10-fold more potent than the parent compound as an antagonist. For the AR transactivation assay, AR() LNCaP prostate carcinoma cells were transfected with an inducible luciferase reporter construct (pGudLuc7ARE) and exposed for 24 h to test materials in the presence and absence of 1 nM of the AR agonist R-1881. Each of these materials, including the hydroxylated metabolites, produced significant antiandrogenic activity in vitro as evidenced by their inhibition of the response to R-1881 (Charles et al., 2005). However, there is no literature at present confirming that this activity is sufficient to induce antiandrogenic effects in vivo (Charles et al., 2005). In the male, exposure to 100 or 200 mg/kg MXC by oral gavage for 70 days damaged Sertoli cells and induced degenerative changes in the spermatogonia and spermatocytes, with some seminiferous tubules devoid of all cellular elements except spermatogonia (Bal, 1984). Using a shorter duration exposure, Linder et al. (1992) administered either 4000 mg/kg for 1 day or 2000 mg/kg for 4 days and reported degenerating cells in stage VII seminiferous tubules 2 days after the acute exposure and similar changes plus remnants of condensed spermatid nuclei in stages VIII–XIV and testicular debris in the caput. Long-term (10-month) exposure of weanling male rats to methoxychlor at levels between 200 and 400 mg/kg/day delayed puberty by as much as 10 days and reduced fertility and copulatory plug formation, sperm counts, and time to pregnancy (Gray et al., 1989). Unlike what was observed for 17-estradiol–implanted rats, no effects of methoxychlor exposure were noted on pituitary weight or on serum LH or prolactin, indicating that the central effects of methoxychlor do not resemble those of endogenous estrogen. The only developmental effect of methoxychlor noted in a standard teratology study in which females were exposed on days 6–15 of gestation to doses of methoxychlor between 100 and 400 mg/kg was an increase in wavy ribs at all dose levels (Khera et al., 1978). However, using other approaches, the finding of heightened sensitivity of the developing organism to estrogens has been confirmed for methoxychlor. For example, mice given 300 mg/kg by oral gavage on gestation days 6–15 were unable to maintain

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pregnancy, while those given 200 mg/kg had prolonged pregnancies and the offspring had an increased percentage of atretic follicles (Swartz et al., 1992). Gray et al. (1989) exposed rats continuously from weaning through puberty and gestation to day 15 of lactation with 25, 50, 100, or 200 mg/kg/day MXC. Treated females displayed accelerated age at vaginal opening and first estrus at all dose levels. Cyclicity was accelerated at 25 mg/kg/day, normal at 50 and 100 mg/kg/day, and abolished at 200 mg/kg/day. Puberty (preputial separation) was delayed in males at the two highest dose levels, whereas growth was reduced at all dose levels. Sex accessory gland weights and caudal sperm counts were reduced at doses as low as 50 mg/kg/day, as were pituitary weights at 100 mg/k/day and above. These findings were extended by Chapin et al. (1997), who exposed female rats to MXC at 0, 5, 50, or 150 mg/ kg/day for the week before and after parturition and then directly dosed the offspring until postnatal day 21. Dosedependent amounts of methoxychlor and metabolites were present in milk and in the plasma of dams and pups. Litter size was reduced at the high dose by 17%. Vaginal opening was accelerated in all dose groups and preputial separation (PPS) was delayed at the two high exposure levels. Adult estrous cyclicity was also altered and fewer ova released at 50 mg/kg and above; all groups of females showed uterine dysplasias, less mammary alveolar development, and reduced estrous FSH levels. In addition, the metabolite HPTE has also been shown to inhibit FSH and cAMPstimulated progesterone production by isolated granulosa cells from immature rats (Zachow and Uzumcu, 2006).

18.2.2  Androgen Receptor Function 18.2.2.1  Linuron Although antiandrogens alter adult male reproductive function, the true impact of their toxicity is not observed until exposures occur encompassing the critical developmental periods when androgens play crucial roles in the differentiation of the reproductive tract and other tissues. Linuron is a chlorinated urea-based herbicide with structural similarity to the nonsteroidal antiandrogen flutamide. It induces Leydig cell adenomas in male rats in chronic bioassays although it lacks genotoxic action in a number of in vitro assays. Cook et al. (1993) exposed adult male rats to 200 mg/kg linuron for 2 weeks to study effects on sex accessory gland weights and the function of the hypothalamic–pituitary axis. Linuron decreased accessory sex gland weights in sexually immature rats and adult treated rats. Increased estradiol and luteinizing hormone (LH) levels were seen in adult treated males. These effects were consistent with the effects of flutamide, although linuron did not elevate serum testosterone as did flutamide. Linuron also competed with [3H]testosterone for binding to the androgen receptor (Lambright et al., 2000; McIntyre,

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2000). The binding affinity was approximately 3.5 times less potent than flutamide in this study. The antiandrogenic effects of developmental exposure to linuron have been described by Gray et al. (1999a) and McIntyre et al. (2000), with induced epididymal and testicular lesions and hypospadias/cleft phallus in male offspring gray. Linuron has also been found to reduce testosterone production ex vivo (Hotchkiss et al., 2004; Wilson et al., 2004).

18.2.2.2  Organophospate Insecticides In 2001, Tamura et al. evaluated the antiandrogenic potential of the organophosphate fenitrothion because of its structural similarities to linuron and the pharmaceutical antiandrogen flutamide. In this study, they demonstrated that fenitrothion competitively antagonized the androgen receptor in transfected HepG2 human hepatoma cells and caused a regression of androgen-dependent tissue weights in the Hershberger rat assay. The inhibition of androgen receptor function in vivo occurred at doses of 15 and 30 mg/kg, while a dose of fenitrothion did not significantly alter blood acetylcholinesterase activity at 15 mg/kg.

18.2.2.3  Dicarboximide Fungicides (a) Vinclozolin One of the first pesticides which demonstrated antiandrogenic effects was vinclozolin, a dicarboximide fungicide used on fruits, vegetables, ornamental plants, and turf grasses. Administration of vinclozolin to adult male rats caused Leydig cell tumors and atrophy of accessory sex glands, including the prostate and seminal vesicles (Van Ravenzwaay, 1992). Gray et al. (1994) reported that developmental exposure of 100 or 200 mg/kg by oral gavage from gestation day 14 to postnatal day 3 in rats resulted in marked demasculinizing effects on male offspring. In both dose groups, male anogenital distance at birth was femalelike, and prominent nipple development was evident at 2 weeks of age. As adults, treated male offspring were unable to attain intromission due to cleft phallus with hypospadias; however, mounting behaviors were normal. Other abnormalities observed included suprainguinal ectopic testes, vaginal pouches, epididymal granulomas, and small to absent sex accessory glands. The only change noted in female offspring was a reduced anogenital distance during the neonatal period. The phenotypic appearance in males is consistent with inhibition of both testosterone-dependent (Wolffian duct differentiation) and dihydrotestosterone-dependent (urogenital sinus and external genitalia) tissues, as expected of an androgen receptor antagonist. This activity was subsequently confirmed by Kelce et al. (1994), who reported that neither vinclozolin nor two principle metabolites [designated Ml (2-[[(3,5-dichlorophenyl)-carbamoyl]oxy]-2-methyl-3-butenoic acid) and

Chapter | 18  Pesticides as Endocrine-Disrupting Chemicals

M2 (3,5-dichloro-2-hydroxy-2-methylbut-3-enanilide)] inhibited 5-reductase activity, but that the metabolites (particularly M2) were able to competitively inhibit binding of [3H]R1881 to the androgen receptor and block androgen-dependent gene activation. Peripubertal exposure of male rats to vinclozolin at doses of 30 and 100 mg/kg/ day retarded sex accessory gland and epididymal growth, but no effects on testes weight or sperm maturation were observed in adulthood (Monosson et al., 1999). A lower dose (10 mg/kg/day) induced significant increases in serum LH and testosterone concentrations. Analysis of serum levels of M1 and M2 suggested that these effects occurred in conjunction with only a low percentage of androgen receptors being occupied, as the levels were below the Ki values from in vitro binding assays. (b) Procymidone Another related dicarboximide fungicide was also shown to bind to the rat and mouse androgen receptor in a study triggered by the observation of hypergonadotropism after 2 weeks of dietary exposure (Hosokawa et al., 1993). Pituitary LH levels were increased after 2 weeks of exposure to 700 ppm in the rat and 5000 ppm in the mouse. Smaller, nonsignificant increases in serum testosterone and LH were noted in both species at the higher exposure concentrations. Scatchard analysis of rat and mouse prostate androgen-receptor binding showed that procymidone had less than 0.07% of the binding affinity of dihydrotestosterone. This affinity was similar to that of flutamide and is sufficient to produce the same spectrum of phenotypes as seen in vinclozolin-exposed male offspring and to inhibit dihydrotestosterone-induced transcriptional activity in CV1 cells cotransfected with the human androgen receptor and a luciferase reporter gene (Gray et al., 1999b; Ostby et al., 1999). In vivo, procymidone appeared to have approximately half the potency of vinclozolin. (c) Iprodione Iprodione (IPRO) is another dichlorophenyl dicarboximide fungicide which also induces Leydig cell tumors in the rat testis following long-term exposures. Although both procymidone and vinclozolin antagonize the androgen receptor (AR) in vitro and in vivo, IPRO does not appear to be an AR antagonist. In weanling rats gavaged with 0, 50, 100, or 200 mg/kg/day from postnatal days 23 to 51, IPRO delayed PPS at 100 and 200 mg/kg/day and decreased androgen-sensitive seminal vesicle and epididymides weights at 200 mg/kg/day (Blystone et al., 2007a). In these animals, serum testosterone and androstenedione were decreased along with ex vivo testis production of testosterone and progesterone. These results suggest that IPRO affects steroidogenesis within the testis, as there was no alteration of serum LH. In addition, IPRO failed to elicit

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an AR antagonism in vitro in a human AR binding assay or the MDA-kb2 reporter gene assay (Blystone et al., 2007a). Therefore, the mechanism of action of IPRO differs from the other dicarboximides, procymidone and vinclozolin.

18.2.3  Steroidogenesis Inhibitors 18.2.3.1  Ketoconazole Ketoconazole, an imidazole antifungal agent first introduced in 1981, is a widely administered oral treatment for systemic mycoses (Fromtling et al., 1998). Its antifungal action is due to its inhibition of the synthesis of ergosterol via the P450dependent enzyme, 14-demethylase (Como and Dismukes, 1994). Ketoconazole is also known to inhibit P450 enzymes of the steroidogenesis system, resulting in adverse endocrine effects in humans (Como and Dismukes, 1994). In clinical studies, ketoconazole was shown to inhibit both adrenal and testicular steroidogenesis (Pont et al., 1982a,b), with short-term decreases in serum androstenedione and testosterone following a single oral dose (De Coster et al., 1985). Ketoconazole inhibits CYP17 and the lyase in the presence of free 17-hydroxyprogesterone. Discovery of these endocrine effects sparked a flurry of interest in other potential clinical applications for this drug. Ketoconazole has been successfully used as a treatment for Cushing’s syndrome and prostate cancer to decrease steroid hormone production (Sonino, 1987). Numerous animal studies have been conducted to explore its potential use as a male contraceptive. Waller et al. (1990) evaluated the effects of ketoconazole on male rat fertility following three consecutive daily oral doses of either 200 or 400 mg/kg/day. Ketoconazole at a dose of 200 mg/kg/day significantly reduced fertility compared to control animals and resulted in a complete loss of fertility at a dose of 400 mg/kg/day. Sperm moti­ lity was reduced at the high dose and forward progression was reduced at both doses. In a similar study with mice, Joshi et al. (1994) also found a significant decline in sperm motility as well as reductions in sperm density at an oral dose of 400 mg/kg administered for a period of 60 days. Fertility in these mice was greatly reduced compared to that of controls. Research has also shown adverse effects on female reproduction. In a study of the effect of ketoconazole on early pregnancy, Cummings et al. (1997) treated rats with 10–100 mg/kg ketoconazole on days 1–8 of pregnancy. Evaluations at gestational day 9 showed a significant reduction in the number of implantation sites and serum progesterone levels as well as increases in uterine body weight. Further test results from pseudopregnant, ovariectomized rats and in vitro ovary culture indicate that ketoconazole directly interferes with uterine function by inhibiting ovarian steroidogenesis. This study confirms earlier research by Buttar et al. (1989), who found intrauterine growth retardation, delayed parturition, and postnatal

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developmental effects such as late descent of testes and vaginal opening in both rats and mice. Using the OECD Enhanced 407, researchers found a reduction in epididymal weights, spermatid retention in the seminiferous tubules, decreased testosterone, and increased estradiol, LH, and FSH. A prolongation of estrous cycles and increased serum estradiol, LH, and FSH in the female were observed (Shin et al., 2006). They also found a decrease in thyroid hormones and an increase in thyroid-stimulating hormone (TSH) but no changes in thyroid weight.

18.2.3.2  Molinate Molinate (S-ethyl hexahydro-1H-azepine-1-carbothioate; CAS No. 2212–67–1) is a thiocarbamate, a preemergent pesticide used on rice fields, particularly in California. Molinate is a thiocarbamate herbicide that has been shown to reduce serum testosterone levels with resulting testicular toxicity (delayed release of spermatids) and impaired fertility in exposed male rats (Minor et al., 1984). In a time course study, Sprague–Dawley rats received a single exposure to 100–400 mg/kg of molinate or 55–200 mg/kg molinate sulfoxide (a major metabolite found in rats) by intraperitoneal injection and were followed for up to 3 weeks. Testicular damage was dose and time dependent following molinate exposure. Histopathological changes (Sertoli cell vacuolation, failed spermiation, and phagocytosis of spermatids at stages X and XI of spermatogenesis) were evident at 2 days after 400 mg/kg and 1 week after 200 mg/kg. With additional time, the lesion progressed until germ cells were virtually absent from the seminiferous tubule. Similar effects were observed with lower doses of the sulfoxide (Jewell et al., 1998). Additional experiments using 14C-labeled molinate, molinate sulfoxide, and molinate sulfone found extensive and tight binding to a protein of 180 kDa, subsequently identified as hydrolase A, a carboxylesterase present in liver and testis (Jewell and Miller, 1998). They hypothesized that inhibition of the esterase could alter the mobilization of cholesterol esters from high-density lipoproteins, thus affecting testosterone biosynthesis. It was subsequently demonstrated that administration of molinate to rats (40–140 mg/kg/day for 7 days) caused a marked decrease in serum and testicular testosterone. In addition, 3H-molinate accumulated in the Leydig cells, and esterase activity in those cells was inhibited. In vitro, molinate sulfonate and molinate sulfone, but not molinate, were potent inhibitors of the esterase activity in Leydig cells (Ellis et al., 1998). A number of regulatory compliance studies describe the potential of molinate to induce toxicity, specifically an adverse effect on reproduction in male rats. Most of these studies remain unpublished but have been summarized in reviews (Cochran et al., 1997; Wickramaratne et al., 1998). In the risk assessment of molinate, it was noted that

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testicular toxicity has not been seen in exposed primates, and epidemiological studies in exposed workers showed no effect, although limitations of those studies did not preclude potential risks to human reproduction (Cochran et al., 1997). In another review, it was noted that spermatoxicity was not seen in molinate-exposed rabbits, dogs, or monkeys, whereas it was in mice and rats (Wickramaratne et al., 1998). The relative order of sulfur oxidation as measured by analysis of urinary metabolites was reported as dog  rat  mousemonkey  rabbit  human. For thiocarbamate cleavage, the rank order was rat  dog   mouse  rabbit  monkey (no data were available for humans). Wickramaratne et al. (1998) argued that the metabolic differences, combined with the unique role of highdensity lipoproteins in cholesterol mobilization in rodents (which is inhibited by metabolites of molinate) as opposed to other mammals which rely on low-density lipoproteins (whose esterase, acetyl-CoA, is not inhibited by molinate metabolites) as their primary source of cholesterol, suggest that the rodent data on testicular toxicity are not relevant to humans. Published data are lacking on the effects of acute exposure to molinate in the female. Recently, molinate was shown to suppress the proestrous LH surge and inhibit the gonadotropin-releasing hormone pulsatility (Stoker et al., 2005). On the day of vaginal proestrus in the female rat, there is a “critical window” for the neural trigger of ovulation. A toxicant which interferes with the hypothalamic regulation of this LH surge will block the LH surge and delay ovulation when administered during this window of time. To examine the effect of molinate on the LH surge, ovariectomized (OVX) rats were implanted with silastic capsules containing estradiol benzoate to mimic physiological levels on proestrus. Single oral doses of 25 and 50 mg/kg molinate significantly suppressed LH and prolactin secretion. Intact regularly cycling females gavaged with 0, 25, or 50 mg/kg molinate at 1300 h on proestrous were examined on estrus or estrus 1 day for the presence of oocytes in the oviduct. All control females had oocytes in the oviduct on estrus. Molinate doses of 6.25 to 50 mg/kg delayed ovulation for 24 h. An extended time in estrous was also shown after daily exposure to 50 mg/kg (21 days). To determine whether molinate blocked the LH surge via a central nervous system (CNS) mode of action or via an alteration in pituitary response, Stoker et al. (2005) evaluated the release of LH in control and molinate-treated rats after a bolus dose of exogenous GnRH. LH release was comparable in the two groups, suggesting that the effect of molinate is centrally mediated. To further examine the potential role of the CNS, they examined the pulsatile release of LH present in the long-term OVX female, in which the pulsatile pattern of LH secretion is directly correlated with GnRH. A significant decrease in the LH pulse frequency was observed in molinate-treated females, thereby indicating that molinate is able to delay ovulation by suppressing the LH surge with

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the brain as the primary target site. Therefore, it appears that molinate or its metabolites have multiple modes of action for disruption of the endocrine system.

18.2.3.3  Triazoles Inhibition of aromatase activity, the enzyme which converts testosterone to estrogen, in the ovarian granulosa cell by the antifungal triazole 1,1di-(4-fluorophenyl)-2-(1,2,4-triazol-1-yl)-ethanol (R151885) has been linked to a blockade of ovulation, particularly when given during diestrus I or II of the estrous cycle (Middleton et al., 1986; Milne et al., 1987). Doses as low as 5 mg/kg by gavage completely suppressed ovulation. When given at midday of diestrus, there were no effects on serum LH, FSH, or progesterone until the afternoon of proestrus. However, plasma estradiol levels were reduced by nearly 50% within 12 h of treatment and remained low for an additional 12 h. The agent was not directly uterotrophic in the ovariectomized mature rat, but doses of 25 mg/kg were able to reduce estradiol-stimulated uterine weight increases.

18.2.3.4  Conazoles (a) Prochloraz Prochloraz is an imidazole fungicide that is widely used in Europe, Australia, Asia, and South America within gardening and agriculture. Screening studies have shown that prochloraz elicits multiple mechanisms of action in vitro, as it antagonizes the androgen and the estrogen receptor, agonizes the Ah receptor, and inhibits aromatase activity. Prochloraz acts as an antiandrogen in vivo (rat Hershberger assay) by reducing weights of reproductive organs, affecting androgen-regulated gene expressions in the prostate, and increasing LH levels. Maternal exposure to prochloraz during the gestational period of sexual differentiation resulted in hypospadias, reduced reproductive organ weights, and increased retention of nipple/areolas in male rat offspring (Laier et al., 2006; Noriega et al., 2005; Vinggaard et al., 2005). Prochloraz was also reported to reduce fetal testosterone production in vivo and ex vivo (Laier et al., 2006; Wilson et al., 2004; Vinggaard et al., 2005). In addition, prochloraz was reported to be an androgen receptor antagonist in vitro and in vivo (Andersen et al., 2002; Noriega et al., 2005; Vinggaard et al., 2002). Prochloraz inhbited CYP 17, which converts progesterone to 17-hydroxyprogesterone, an intermediate which can disassociate from the enzyme or be further converted to androstenedione through CYP 17 lyase activity. Prochloraz inhibited the hydroxylase activity in a similar fashion to other imidazoles (Ayub and Levell, 1987). Prochloraz may also inhibit the CYP 17 lyase, as Blystone et al. (2007b) found an increase in 17-hydroxyprogesterone.

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18.2.4  Inducers of Steroid Clearance The pregnane X-receptor (PXR) and the constitutive androstane receptor (CAR) are orphan nuclear receptors activated by a variety of ligands (Kretschmer and Baldwin, 2005). These receptors may act as xenobiotic/steroid sensors and regulators of chemical stressors. They are both important regulators of several steroid and xenobiotic detoxification enzymes and transporters in the liver and intestine. The proteins that they induce in detoxification include metabolism, deactivation, and transport of bile acids, thyroid and steroid hormones, numerous environmental chemicals, and several pharmaceuticals. Interestingly, some steroids and steroid mimics activate either one or both of these receptors, including several endocrine disrupting chemicals. Environmental estrogens including methoxychlor, endosulfan, dieldrin, and DDT activate either PXR or both receptors. PXR and CAR induce enzymes, such as CYP2B and CYP3A family members, responsible for the metabolism of steroid and thyroid hormones, thus altering normal physiological function. For example, chlordane is an agonist of the human and rodent CAR, as is lindane. Methoxychlor is an agonist of PXR and CAR and vinclozolin is an agonist of the rodent PXR (for review, see Kretschmer and Baldwin, 2005).

18.2.5  Enhancers of Steroid Action 18.2.5.1  Triclocarban and Triclosan Triclocarban (TCC; 3,4,4-trichlorocarbanilide), an antimicrobial compound, is commonly added to a wide range of household and personal care products including bar soaps, detergents, body washes, cleansing lotions, and wipes for its sanitizing properties. Triclocarban-containing products have been marketed broadly for more than 45 years and thus have a long history of use in Europe and the United States. It is estimated that approximately 1 million pounds are produced for the U.S. market per year, and recent reports suggest widespread contamination of U.S. water resources. A study found that TCC does not compete with the endogenous hormone for receptor binding but amplifies the androgen receptor-mediated, native androgen-induced transcriptional activity in vitro and in vivo (Chen et al., 2008). Another antimicrobial which is structurally similar to triclocarban is triclosan [5-chloro-2-(2,4-dichlorophenoxy)phenol] and it has also been shown to potentiate or amplify the estrogen-mediated response in the weanling rat uterotrophic assay (Stoker et al., 2009). Although no response was observed with triclosan alone, there was a dose–response increase in uterine weight and morphology when triclosan was co-administered with ethinyl estradiol. These types of steroid enhancers may be acting through several different mechanisms, such as interference with steroid clearance or by acting as coactivators of the receptor–protein complex upon binding of the steroid.

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18.3  Hypothalamic–pituitary– gonadal axis 18.3.1  Dithiocarbamates Dithiocarbamates are a broad chemical class including fungicides such as the ethylenbisdithiocarbamates, metam sodium and thiram. They are also metal chelating agents and are known to inhibit the synthesis of neurotransmitters, particularly norepinephrine, via chelation of the coppercontaining portion of the enzyme dopamine--hydroxylase. Norepinephrine plays a critical role in the release of gonadotropin-releasing hormones (GnRH) from the hypothalamus. During a short time period (between 1400 and 1600 h on the day of proestrus), the sequential feedback of estrogen and then progesterone stimulates the activity of -adrenergic neurons which induces a phasic release of GnRH. This, in turn, triggers the anterior pituitary to release a surge of LH. Concentrations of LH rapidly rise in serum from levels below 1 ng/ml to 5–10 ng/ml and ovulation is induced. There is a minimum concentration for this surge to be effective in inducing ovulation in spontaneous ovulators such as the rat and the human. Because of the critical timing of events, and the multiple steps which are susceptible to disruption, analysis of the control of ovulation has proven a particularly useful tool in understanding neuroendocrine toxicology. Ovariectomized, estrogen-primed female rats given a single injection of 50 or 100 mg/kg thiram at 1100 h suppressed the estrogen-induced LH surge (Stoker et al., 1993). Administration of 50 mg/kg to intact females on the afternoon of vaginal proestrus delayed ovulation by 24 h. When mated the following evening, there were no differences in the proportion of sperm-positive females compared to controls mated the previous evening, but there was a significant decrease in embryo viability between gestation days 7 and 11, indicating that ovulation delayed for 24 h was deleterious to the ripening oocytes (Stoker et al., 1996), which was later shown to be due to altered cortical granule distribution and polyspermic zygotes (Stoker et al., 2003).

18.3.2  Atrazine The herbicide atrazine is widely used in agriculture as a pre-emergence herbicide for corn and other crops. Because of its physical and chemical properties, atrazine is found in small concentrations in surface waters. Concern for the endocrine-disrupting effects of this chlorotriazine herbicide arose following the initial observation of increased incidence and earlier onset of mammary tumors in a chronic bioassay (over 100 days) in female Sprague–Dawley (SD) rats exposed to 400 ppm atrazine in the diet (Stevens et al., 1994; Thakur et al., 1998). The finding of an earlier onset of mammary tumors led to an investigation into the estrogenicity of atrazine by a

Hayes’ Handbook of Pesticide Toxicology

number of investigators. However, under equilibrium conditions, atrazine was not able to compete with estradiol for binding to rat uterine estrogen receptors. A weak competition was noted if the cytosols were preincubated at 25°C prior to incubation with the tracer (Tennant et al., 1994a). Somewhat conflicting results have been seen in other studies. Daily exposure of adult Fischer rats to 120 mg/kg for 7 days resulted in fewer treated females displaying normal estrous cycles, and the number of days in diestrus increased significantly. Fertility was reduced in females during the first week after exposure, but pregnancy outcome was not affected in those that became inseminated (Simic et al., 1994). However, treatment of adult, ovariectomized SD rats with up to 300 mg/kg atrazine by oral gavage for 3 days did not result in an increase in uterine weight, nor were there increases in uterine progesterone levels, suggesting the lack of an estrogenic potential. Indeed, when estradiol (2 g/kg subcutaneously) was given in conjunction with 300 mg/kg of orally administered atrazine, there was a weak inhibition (25%) of the uterotrophic response (Tennant et al., 1994b). In a similar study, immature female SD rats were dosed with 0, 50, 150, or 300 mg/kg atrazine by gavage for 3 days. Uterine weight was not increased, but decreases in uterine progesterone receptors and peroxidase activities were noted; however, when combined with estradiol, antiestrogenic effects of atrazine including decreases in uterine progesterone receptor binding and uterine peroxidase were not noted on the uterus (Connor et al., 1996). In this same study, atrazine did not affect basal or estradiol-induced MCF-7 cell proliferation, nor did it display agonist or antagonist action against estradiol-induced luciferase activity in MCF-7 cells transfected with a Gal4-regulated human estrogen receptor chimera. To further evaluate effects on reproductive function, female Long Evans (LE) and SD rats that had been screened for regular 4-day estrous cycles received 0, 75, 150, or 300 mg/kg/day atrazine by gavage for 21 days. In both strains, atrazine disrupted the regular 4-day estrous cycles. For the LE rats, all dose levels were effective, whereas SD rats required a higher dose (150 mg/kg/day) for a longer time for this effect to appear. The increased time spent in vaginal diestrus was associated with elevated serum progesterone and low estradiol concentrations, indicative of a repetitive pseudopregnant condition. This hormonal condition was not considered to be conducive to the development of mammary tumors, although there was some indication of prolonged estrous at the lowest dose tested (Cooper et al., 1996). Reproductive cycling in the female SD rat begins to decline in animals less than 1 year of age, presumably due to the loss of sensitivity of adrenergic neurons in the hypothalamus that control GnRH release to the pituitary. This loss of stimulation reduces FSH and LH release, and ultimately ovulatory failure. In turn, the ovaries contain

Chapter | 18  Pesticides as Endocrine-Disrupting Chemicals

many follicles but no corpora lutea. Hence, the endocrine milieu of the aging SD rat favors development of mammary tumors. How atrazine accelerates the neuroendocrine aging of the reproductive axis in the SD rat, however, has not been determined. Several studies (Laws et al., 2000b; Stoker et al., 2000) have evaluated the effects of atrazine on pubertal development in the male and female Wistar rat. Atrazine exposure delayed the onset of vaginal opening and altered estrous cyclicity in the female rats following oral exposure from weaning to post-puberty (PND 22 to 41). The lowest-observable-adverse-effect level (LOAEL) for vaginal opening and subsequent estrous cyclicity was 50 mg/kg, and the no-observable-adverse-effect level (NOAEL) was 25 mg/kg. The effect on the estrous cycle was reversible as shown by the resumption of normal estrous cyclicity by all females by 30 days following the last exposure to atrazine. This data was consistent with the effect of atrazine on the CNS and subsequent alterations in the hormonal control of pubertal development. Others have shown similar results of atrazine on female pubertal development with other strains of female rats. The NOAEL was 30 mg/kg in the Aderley Park strain (compared to 25 mg/kg in the Laws et al. study) and 10 mg/kg in the SD strain for delayed vaginal opening (VO) (Ashby et al., 2002). These investigators also compared the effect of atrazine to a dose of the GnRH antagonist, Antarelix, and found similar delays in VO in the Aderley Park strain, as a comparison of the proposed mechanism of action of disrupted GnRH signaling during the peripubertal period. In the male rat, there were delays in the onset of puberty following exposure from PND 23 to 53, as determined by assessment of preputial separation. Stoker et al. (2000) found that peripubertal exposure to atrazine delayed puberty for 1.5–2.5 days following exposure to 12.5–200 mg/kg. The NOAEL for the effect on puberty in the male Wistar rat was 6.25 mg/kg/day and the LOAEL was 12.5 mg/kg. They also observed significant decreases in the growth of the reproductive tract, namely the seminal vesicles and prostates, of the affected male rats at slightly higher doses than the doses required to delay puberty (NOAEL  25 mg/kg). In the studies examining the SD rat (Trentacoste et al., 2001), pubertal onset was delayed by 3 and 4 days following exposure to 100 and 200 mg/kg atrazine, respectively, with decreases in ventral prostate and seminal vesicles at the same doses (NOAEL  50 mg/kg). All three studies found decreases in either serum or intratesticular testosterone, and Trentacoste et al. (2001) also found a decrease in LH at the 200 mg/kg dose. The effects on pubertal progression in the male and the delay of reproductive tract growth agree with the mode of operation (MOA) of a decreased secretion of LH. For example, ventral prostate, seminal vesicles, lateral prostate, and epididymis weights were all significantly reduced at doses just above the doses which delayed

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preputial separation (Stoker et al., 2000). These effects provide evidence of a decrease in androgen stimulation. Decreases in androgen hormone concentrations also correspond with the effects observed on pubertal endpoints. For example, Stoker et al. (2000) demonstrated that intratesticular testosterone was significantly decreased following atrazine exposure peripubertally. This has corresponded to the work of others which has also shown that serum testosterone is decreased by doses of 50, 100, and 200 mg/kg in a different strain of rat (SD) when dosed during a similar period of PND 22–47 (Friedmann et al., 2002; Hayes et al., 2006). A number of reports on the effects of atrazine on aquatic vertebrates, mostly amphibians, have been published, yet there is inconsistency in the effects reported and inconsistency between studies in different laboratories (Hayes et al., 2006; Kloas et al., 2009a,b). These studies examined the growth, larval development, or sexual differentiation in Xenopus laevis tadpoles following exposure to 0.01–100 g/l during juvenile development. Based on a weight of evidence analysis of all of the data, the central theory that environmentally relevant concentrations of atrazine affect reproduction and/or reproductive development in fish, amphibians, and reptiles is not supported by the vast majority of observations. Whether or not the same conclusions would also hold for the hypothesis that atrazine enhances aromatase activity cannot be determined until the relevant measures of this enzyme are measured in the different species.

18.3.3  Pyrethroids Pyrethroid insecticides have been used in agricultural and home formulations for more than 30 years and approximately 16 are registered for use in the United States (Bryant et al., 2003). These synthetic chemicals are typically sold or used as mixtures containing two or more insecticides. The toxic effects of pyrethroids are caused by the prolongation of the open state of voltage-dependent sodium channels, resulting in repetitive firing of the neurons. High-dose in vitro studies have shown the antiandrogenic activity of several pyrethroids (Eil and Nisula, 1990) with Km values in the micromolar range. Some have been found to be estrogenic in vitro (Garey and Wolff, 1998; Go et al., 1999; Kim et al., 2004b). Cypermethrin was given to pregnant rats by subcutaneous injections at dose levels between 0.25 and 25 mg/kg/day for the last 7 days of gestation, and pups continued to be dosed postnatally until 30 days of age. Anogenital distance was reduced at birth and at 85 days of age in males. There was a dose–response decrease in relative prostate weight at 55, but not 85, days of age, and there were no effects on epididymal sperm counts at that time (Ronis and Badger, 1995). These results suggest a weak antiandrogenic effect of the chemical.

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Permethrin, a synthetic pyrethroid insecticide, has been shown to alter endocrine function. In the uterotrophic assay using 18-day-old female rats, permethrin (200 and 800 mg/kg) s.q. for 3 days increased relative uterine wet weights and E2-induced uterine weights. This effect was inhibited by the co-administration of ICI 182,780, an antiestrogen. Another study examined the effects of permethrin on uterine Calbindin-D9k (CaBP-9k) gene expression and in a uterotrophic assay. The CaBP-9k gene, one of the intracellular calcium-binding proteins, is estrogen responsive in the uterus. Northern blot analysis showed the induction of uterine CaBP-9k mRNA level in response to permethrin as well as co-administration of permethrin with E2. In the Hershberger assay, the administration of permethrin orally to testosterone propionate-treated castrated male rats led to statistically significant reductions in androgen-dependent sex accessory tissue (ventral prostate, seminal vesicles, levator ani and bulbocavernosus muscles, Cowper’s gland, and glans penis) weights at all doses tested (10, 50, and 100 mg/kg). These results suggest that permethrin might have estrogen-like effects on female rats but antiandrogenlike effects in males (Kim et al., 2005). Although there have been some reports that esfenvalerate possesses antiestrogenic properties (Kim et al., 2004a), in vivo screening tests revealed no effects in the Hershberger and uterotrophic assays (Kunimatsu et al., 2002). One report found that a short-term exposure to 0.5 or 1.0 mg/kg esfenvalerate delays the onset of puberty in the female rat by altering the hypothalamic control of prepubertal LH secretion (Pine et al., 2008). Lastly, lambda-cyhalothrin is used to control mosquitoes, ants, cockroaches, etc. in households and has been shown to be estrogenic in vitro by inducing the proliferation of MCF-7 cells, which were blocked by the ICI 182,780 anti-estrogen. It was also shown to decrease the production of mRNA for both ERa and ERb.

18.4  Thyroid hormone Endocrine disruption of the pituitary–thyroid axis is a relatively well understood process by which endogenous chemicals induce thyroid follicular cell neoplasia. The physiological regulation of thyroid cell growth and function involves a complex interactive network of trophic factors that are mediated by a number of second messenger systems (Hard, 1998). TSH is the main growth factor for follicular cells, with insulin-like growth factor 1, epidermal growth factor, basic fibroblast growth factor, and transforming growth factor-beta also involved in various ways. Activation of TSH receptors stimulates G protein-dependent elevation of cAMP and phospholipase C, which in turn regulates iodine uptake and release, thyroid peroxidase (TPO) generation, thyroid hormone synthesis and release, and thyroid cell growth and division. Relative to metabolism,

T4 is secreted by the thyroid but must be converted to T3 via either type I 5-diodinase in the liver or type II 5-diodenase in the brain, pituitary, and brown adipose tissue. There are three main carrier proteins for thyroid hormones – thyroxine binding protein (65%), transthyretin (20%), and albumin (10%); only about 5% of the hormone is unbound (in the rat, thyroxine binding protein is absent during most of adult life). Further metabolism occurs in the liver, intestines, and kidneys and involves inactivation of biological activity by conjugation with glucuronic acid or sulfate. Whether by reduced synthesis due to inhibition of TPO, reduced peripheral de-iodination, or elevated turnover via induction of conjugating enzymes, sustained release of TSH in response to decreased circulating levels of thyroid hormones is intimately involved in thyroid gland neoplasia. This suggests that nonlinear thyroid cancer dose–response considerations can be applied to chemicals that reduce thyroid hormone levels, increase TSH and thyroid cell division, and are judged to lack mutagenic activity (Hill et al., 1998). Although much of thyroid gland physiology is similar across experimental animals and humans, there are, as noted previously, some important differences that may reduce the sensitivity of humans relative to rodents (Hard, 1998). Interestingly, childhood radiation is the only known exogenous risk for thyroid gland carcinogenesis in humans (Robison, 2009). Several pesticides (i.e., amitrole, ethylene thiourea, and mancozeb) have been shown to induce a high incidence of thyroid tumors (0.48) at relatively low daily doses (3.5– 30.9 mg/kg/day). There are several potential sites of antithyroid action by pesticides, including inhibition of iodide uptake, inhibition of thyroid peroxidase, damage to thyroid follicular cells, inhibition of thyroid hormone release, inhibition of 5-monodeiodenase activity, and enhancement of metabolism and excretion by the liver. Amitrole, ethylene thiourea, and mancozeb inhibit thyroid peroxidase; amitrole, ethiozin, ethylene thiourea, and pentachloronitrobenzene inhibit the iodide pump; 2,4-D, acetochlor, clofentezine, fenbuconazole, fipronil, pendimethrin, pentachloronitrobenzene, prodiamine, pyrimethanil, and thiazopyr stimulate thyroid hormone metabolism and excretion. Ketoconazole was shown to decrease the serum concentrations of thyroxine and triiodothyronine and increase TSH in both the male and the female in the 28-day repeated dose toxicity study (OECD Enhanced TG407) (Shin et al., 2006). Some of the conazoles have been shown to induce thyroid tumors in rats (Hurley, 1998) due to the upregulation of a number of P450 enzymes in the liver, including ketoconazole (Ronis et al., 1994).

18.4.1  2,4-D 2,4-D (2,4-dichlorophenoxyacetic acid) is a herbicide in the phenoxy or phenoxyacetic acid family that is used

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postemergence for selective control of broadleaf weeds. Residents and professional applicators may use 2,4-D on home lawns. At concentrations that exceed the dose for renal clearance following chronic exposure, it has been shown to cause hypothyroid state and thyroid tumors in rats. In addition, there is a significant suppression of thyroid hormone levels in ewes dosed with this chemical (Rawlings et al., 1998). Similar findings have been reported in rodents, with suppression of thyroid hormone levels, increases in thyroid gland weight, and decreases in weight of the ovaries and testes (Charles et al., 1996). The increases in thyroid gland weight are consistent with the suppression of thyroid hormones since the gland generally hypertrophies in an attempt to compensate for insufficient circulating levels of thyroid hormones. Thyroid hormone is known to play a critical role in the development of the brain. Slight thyroid suppression has been shown to adversely affect neurological development in the fetus, resulting in lasting effects on child learning and behavior (Haddow et al., 1999). 2,4-D causes slight decreases in testosterone release and significant increases in estrogen release from testicular cells (Liu et al., 1996). In rodents, this chemical also increases levels of the hormones progesterone and prolactin, and it causes abnormalities in the estrus cycle (Sturtz et al., 2008). Male farm sprayers exposed to 2,4-D had lower sperm counts and more spermatic abnormalities compared to men who were not exposed to this chemical (Lerda and Rizzi, 1991). 2,4-D also interferes with the neurotransmitters serotonin and dopamine. In young organisms, exposure to 2,4-D results in delays in brain development and abnormal behavior patterns, including apathy, decreased social interactions, repetitive movements, tremor, and immobility (Evangelista de Duffard et al., 1995). Females are more severely affected than males. Rodent studies have revealed a region-specific neurotoxic effect on the basal ganglia of the brain, resulting in an array of effects on critical neurotransmitters and adverse effects on behavior (Bortolozzi et al., 2001). In addition, another study showed that 2,4-D had an estrogen response in breast cancer cells (Lin and Garry, 2000). These investigators also evaluated the cellular and molecular developmental toxicity of 2,4-D and other pesticides commonly used in Red River Valley, Minnesota (Lin and Garry, 2000). In addition, a study from The Netherlands showed that 2,4-D has the ability to displace sex hormones from the protein that normally transports these hormones in the blood (Meulenberg, 2002).

18.4.2  Triclosan Another similar in structure to 2,4-D is triclosan (5-chloro2-(2,4-dichlorophenoxy)phenol), an antibacterial compound found in many household soaps and personal care products, as well as used in infant toys. It is used as an active ingredient in many personal care and household products,

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including soaps and toothpaste, and it is detected routinely in wastewater effluents, aquatic species, and human breast milk. Previous research had implicated triclosan as an endocrine disrupting chemical, with effects on thyroid hormone and estrogen/androgen activity (see Section 18.2). In one study, weanling male rats were exposed to triclosan by oral gavage for 31 days and examined for effects on pubertal development and thyroid function (Zorrilla et al., 2009). The doses administered were 0–300 mg/kg/day by oral gavage. The triclosan exposure resulted in a dosedependent decrease in thyroxine (lowest-observable-effect level of 30 mg/kg), a decrease in T3 (200 mg/kg only), increased liver weights (100 mg/kg and higher), and an increase in the hepatic enzyme activity (7-pentaoxyresorufin O-pentaylase or PROD). These data suggested that triclosan may increase the metabolism of thyroid hormone through the induction of hepatic enzymes. No effect was observed on any of the pubertal endpoints, such as preputial separation and androgen-dependent tissue weights.

18.5  Impact on testing guidelines 18.5.1  Multigenerational Studies Assessment of the potential developmental and reproductive risks of environmental contaminants is generally determined through application of testing guidelines that are established by regulatory agencies such as the U.S. EPA with combined efforts with the international coordinating bodies, such as the Organization of Economic Cooperation and Development (OECD). The multigeneration studies were designed to identify developmental and reproductive effects by examining parental animals and offspring dosed pre- and postnatally to establish a NOAEL for the most sensitive effects, thus providing the basis for quantitative assessments. Traditionally, these tests had been very apical in nature; that is, they relied on endpoints which were diagnostic of adverse biological outcomes but did not provide clarification of potential modes of action, target organs, or most sensitive life stage or gender. For example, the multigeneration reproductive test guidelines require groups of animals (generally rats) to be exposed to the test chemical beginning shortly after weaning and continuing until production of the second generation (thus concluding with examination of offspring of animals exposed from fertilization and through reproduction). In the Pre1998 Multigenerational Studies, the primary endpoints that were evaluated included fertility (Are the animals capable of reproducing?), fecundity (How many offspring are produced?), and growth of the offspring. A large body of evidence suggested, especially for endocrine-disrupting chemicals, that these endpoints are neither very sensitive to reproductive disturbance nor indicative of the underlying biological effect. Both these issues raised concern regarding

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the suitability of previously issued test guidelines to satisfactorily detect and characterize reproductive hazard. A review of the data from multigeneration studies indicated that those chemicals that act via the estrogen receptor (e.g., methoxychlor), the androgen receptor (e.g., vinclozolin), or the Ah receptor (e.g., dioxin) can be identified by the traditional endpoints of fertility, fecundity, and growth do identify reproductive toxicants. However, such measures are less sensitive than specific endocrine-dependent endpoints such as vaginal opening (to detect estrogenic compounds) and preputial separation (to detect antiandrogens), indicating that such measures may influence the overall NOAEL or LOAEL of the compound. For example, one of the most sensitive indicators of developmental exposure to an estrogen is accelerated puberty in the female (Gray et al., 1988), whereas diminished anogenital distance and accessory sex gland weights are most sensitive to developmental exposure to an antiandrogen, and decreased ejaculated sperm counts are the most sensitive to chemicals that act via the Ah receptor (Gray et al., 1995). In 1998, the U.S. EPA published a harmonized multigenerational testing guideline in order to improve the ability to detect the effects of chemicals that may act via the endocrine system to perturb reproduction. The guideline (U.S. EPA, 1998a) was updated to include a number of endpoints that monitor reproductive performance and health. These included assessments of the following: female estrous cyclicity; sperm parameters (total number, percentage progressively motile, and sperm morphology in both the parental and F1 generations); the age at puberty in the F1 generation (vaginal opening in the female and preputial separation in the males); an expanded list of organs for pathology, gravimetric analysis, and/or histopathology to identify and characterize effects at the target organ; and some triggered endpoints including anogenital distance in the F2 generation and primordial follicular counts in the parental and F1 generations. For the prenatal developmental toxicity test guidelines (U.S. EPA, 1998a), one important modification related to the improved detection of endocrine disruptors was the expansion of the period of dosing from the end of organogenesis (i.e., palatal closure) to the end of pregnancy in order to include the developmental period of urogenital differentiation. Collectively these modifications of the test guidelines markedly improved the characterization of endocrine-mediated effects during reproduction and development. Although the post-1998 protocol has been useful, there were still concerns as to whether or not improvements could be incorporated to make such testing even more efficient and relevant to the risk assessment processes. In 2006, the Agricultural Chemical Safety Assessment (ACSA) Technical Committee of the International Life Sciences Institute (ILSI) Health and Environmental Sciences Institute (HESI) published a series of articles addressing

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these needs. With input from several international groups, a new test called the Extended One Generation Reproductive Toxicity Test is now under consideration as a possible test guideline (Cooper et al., 2006). This protocol uses a flexible approach that encourages the use of toxicokinetics (ADME – absorption, disposition, metabolism, and elimination) for dose setting, evaluates more pups per litter in the F1 (3 vs. 1), and only requires mating of the F1 (to produce F2) if triggered by certain effects in the PO and developing F1 rats. The additional pups per litter provide for more statistical power and the additional endpoints allow for better diagnostics for interpreting the mechanism of action (Table 18.1). The inclusion of a Neurotoxicity and Immunotoxicity cohort was also added (Table 18.1). Compared to the existing protocol, this proposed protocol would use fewer animals, provide additional information on the developing F1 animal, and include an estimation of human exposure potential to help with decisions on extent of testing required. These efforts took place between 2006 and 2009 by a panel of government and industrial scientists (convened in 2007) and through discussions with an OECD expert group. Therefore, the newly suggested protocol offers a great degree of flexibility while maintaining the scientific rigor needed in a guideline study. Eliminating the need for automatic progression to the production of an F2 appears to have gained credibility through various retrospective analyses that have been conducted (Janer et al., 2008; Reaves et al., 2008) and would have cost savings benefits. The other gain is the opportunity to examine neurotoxic and immunotoxic effects without increasing the number of animals produced in the F1.

18.5.2  Endocrine Disruptors Screening Program In the United States the mandate for testing for EDCs was formalized with the passage of the Food Quality Protection Act (FQPA) in 1996, and a subsequent amendment to the Safe Water Drinking Act (SWDA, 42[(Page 56450)] U.S.C 300j-17) (U.S. EPA, 2009), which required the U.S. EPA to develop a screening program, using appropriate validated test systems and other scientifically relevant information, to determine whether certain substances may have an effect in humans that is similar to an effect produced by a naturally occurring estrogen, or other such endocrine effect as the Administrator may designate [408(p), FFDCA, 21 U.S.C. 346a(p)].

The agency responded by forming an Endocrine Disruptor Screening Program that created and consulted multiple stakeholder committees of scientist experts. Although the congressional mandate only required identification of contaminant EDCs with estrogen-related activities

Chapter | 18  Pesticides as Endocrine-Disrupting Chemicals

Table 18.1  Evolution of Measurements Indicative of Altered Endocrine Function in Multigenerational Guidelines Pre-1998 Post-1998 Ext. One-Gena Fertility

Yes

Yes

Yes

Fecundity

Yes

Yes

Yes

Growth of F1

Yes

Yes

Yes

No. of F1 maintained to adulthood

1

1

3

AGD (sex. differentiation)

No

No (2nd gen)

Yes (F1)

Nipple retention (F1 male)

No

No

Yes

Puberty

No

Yes

Yes

Estrous cyclicity

No

Yes

Yes

Gamete number

No

Yes

Yesb

Thyroid homones

No

No

Yes

F1 mating

Yes

Yes

Triggered

Neurotoxicity (F1)

No

No

Yes

Immunotoxicity (F1)

No

No

Yes

No

No

Yes

c

Use of ADME data for design

b

a

Proposed. Includes both sperm parameters and follicle counts. c Absorption, disposition, metabolism, and elimination, obtained in a satellite study. b

in humans, the agency moved forward with the recommendation of their premier advisory committee [Endocrine Disruptor Screening and Testing Advisory Committee (EDSTAC)], to include androgen and thyroid hormone pathways, and also the hypothalamic–pituitary–gonadal axis, as targets for screening and testing. In addition, EDSTAC recommended that wildlife models of endocrine disruption be added to the screening and testing program (U.S. EPA, 1998b).

18.5.2.1  The Two-tier Concept The task of developing a strategy to identify endocrine disruptors was difficult, as there are a wide variety of hormones, endocrine functions, and mechanisms that could be targeted by an environmental contaminant. The focus was first limited to estrogen, androgen, and thyroid (EAT) hormones and later the hypothalamic–gonadal hormones. The first advisory committee, EDSTAC, produced a recommendation for a “two-tier” approach for screening and testing.

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Tier 1 assays would provide a more qualitative “yes-or-no” screen of tested substances, with chemicals which tested positive undergoing further evaluation with tier 2 tests, to provide dose–response relationships, confirm mechanisms of action, and determine adverse effects in multiple species. With large numbers of potential endocrine disrupting candidates in the environment, it was recommended that the agency adopt a screening approach that would efficiently distinguish those candidates with endocrine disrupting properties. Tier 2 assays would be in vivo, covering several classes of vertebrates and invertebrates and including exposure through various routes during reproduction. The tier 1 battery included a number of in vitro and in vivo assays for redundancy of endpoints. This would help regulators conduct a “weight-of-evidence” evaluation based on a mechanism of action approach to prioritize and sort the results. For example, concluding that a substance is estrogenic would only be made on the basis of several estrogendependent results from the tier 1 screening (TIS) battery. The main attributes of the TIS battery are (1) enhanced sensitivity (e.g., in vitro assays); (2) capacity to minimize false negatives (e.g., in vivo assays); (3) diversity to cover multiple modes of action (e.g., receptor- and nonreceptormediated effects); and (4) range to address effects in mammalian and nonmammalian species within and between gender (e.g., rodents, amphibians, and fish). The main attributes of the tier 2 tests include (1) identify hazard; (2) quantitative relationships between dose and adverse effects; and (3) provide information needed to make a comprehensive assessment of risk. Hence, only after tier 2 testing will the U.S. EPA be able to determine whether a particular chemical substance may have an effect on humans and wildlife that is similar to the effect produced by naturally occurring EAT hormones and be considered an endocrine disruptor. FQPA requires that the U.S. EPA test all pesticide chemicals, including all active and inert ingredients, for endocrine disrupting activity. The agency was also granted the authority to provide for the testing of any other substance that may have an effect that is cumulative to an effect of a pesticide chemical if the Administrator determines that a substantial population may be exposed to such a substance (21 U.S.C., 346a(p).

Also the amendment to the SDWA adds the authority to test any substance that may be found in sources of drinking water (Public Law 104–182, 104th Congress) (U.S. EPA, 2009a). Under the Toxic Substances Control Act (TSCA) and Federal Insecticide, Fungicide, and Rodenticides Act, the U.S. EPA also retains the authority to require testing of industrial chemicals or pesticides undergoing registration review if determined that the chemical may have endocrine disrupting effects (U.S. EPA, 2009b). Given the vast

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number of chemicals covered under these legal regulations, the agency has developed and published an approach for prioritizing and selecting chemicals to be tested in the EDSP and followed this with a list of 67 chemicals that likely will be included in the first test orders for the EDSP (U.S. EPA, 2009c). The draft list (including over 60 active pesticides and some high production volume inert pesticides) was selected based upon the potential for human exposure. This potential was determined by consumption by food or drinking water containing pesticide residues, residential use of pesticide products, volume of use, and occupational contact with pesticides. Selection of a chemical for this list does not mean that it is a known or suspected endocrine disruptor in humans or other species. The agency will submit data from the first set of chemicals tested in the EDSP to an independent expert panel that will be asked to review the results and provide recommendations for improvement of the TIS battery. The list of 67 chemicals can be viewed at an EDSP website (U.S. EPA, 2009d).

The current TIS battery (website with battery; Table 18.2) is the product of a difficult series of scientific evaluations and deliberations by the U.S. EPA’s EDSP and multiple advisory committees (e.g., EDSTAC, EDMVS, and EDMVAC), the Interagency Coordinating Committee on the Validation of Alternative Methods, and SAP reviews. The general public, stakeholders, and the regulated community have participated in open Federal Advisory committee meetings and workshops, as well as in public comment periods advertised in the Federal Register, and have provided input and feedback on the development of the assays for the battery. The collaboration between the U.S. EPA and the OECD has promoted the international harmonization of a five-stage validation process, a joint effort to develop and validate the assays in the TIS battery, and the development of test guidelines for assays that can be used for the screening of chemicals for endocrine disrupting activity. The performance and optimization of the assays will be evaluated by an external panel of experts following the

Table 18.2  Assays Being Proposed for the Tier 1 Screening Battery and Mechanisms Detected by Each Estrogen

Androgen

Estrogen and androgen

Thyroid

Assay E

Anti-E

ER binding (rat uterus)

X

X

ER transcriptional activation (human)

X

A

Anti-A

X

X

Steroidogenesis

HPG

HPT

X

X

X

X

X

X

X

X

In vitro

AR binding (rat prostate) Steroidogenesis H295R (human)

X

Aromatase recombinant (human)

X

In vivo Uterotrophic (rat)

X

Hershberger (rat)

X

X

Pubertal male (rat)

X

X

Pubertal female (rat)

X

X

Fish screen (short-term repro)

X

X

Amphibian metamorphosis (frog)

X

X

X

Chapter | 18  Pesticides as Endocrine-Disrupting Chemicals

first set of test chemicals and as the process moves forward. New methods and state-of-the-art approaches are continuously developed and these new techniques will be considered for inclusion. In addition, studies are ongoing to complete the validation of a number of stably transfected cell lines for use in the AR transcriptional activation assay, as well as recombinant rat and chimpanzee AR binding systems. This will eliminate the need for animals currently consumed by the AR cytosolic binding assay. The agency is also considering the use of performancebased test guidelines (PBTG) as a way to allow substitutions for a number of assays that basically have the same function as those in the TIS battery. For this purpose, the OECD is developing a draft PBTG for the ER transcriptional activation assay (HeLa 9903 cell line) that could ultimately allow ER TA assays with newer technology and/or proprietary components to be used. New information obtained from emerging technologies in the areas of genomics, proteomics, metabolomics, and computational toxicology will also be incorporated into the second-generation TIS battery. The U.S. EPA’s National Center for Computational Toxicology is developing databases using state-of-the-art high-throughput screening bioassays of chemicals under their ToxCast program (http://www.epa.gov/ncct/toxcast) and data generated by the TOX21 initiative (http://www.alttox.org/ttrc/overarching-challenges/way-forward/austin-kavlock-tice/initiative) might play a significant role in the priorization and selection of chemicals to be tested in the next phase of the EDSP. Thus, the TIS battery will continue to be refined and updated as the agency’s EDSP evolves toward full implementation.

Conclusion Pesticides are designed to be bioactive against certain targets but can cause toxicity to nontarget species by a variety of other modes of action including disturbance of endocrine function. As such, pesticides have been found to bind and alter the function of hormone receptors, alter the synthesis or clearance of endogenous hormones, interact with various neurotransmitter systems, and cause yet other effects by still poorly understood mechanisms. The pesticides which produce these effects on the endocrine system encompass a variety of pesticide chemical classes. Some of these pesticides are pervasive and widely dispersed in the environment. Some are persistent, can be transported long distances, and others are rapidly degraded in the environment or the human body. However, even a brief exposure to pesticides which alter endocrine function can cause permanent effects if the exposure occurs during critical windows of reproductive development.

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Acknowledgments The authors gratefully acknowledge the help of Dr. Leah Zorrilla (NCSU College of Veterinary Medicine/U.S. EPA Coop) and Dr. Susan Laws (Endocrinology Branch, U.S. EPA) with the preparation and review of this chapter.

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Garey, J., and Wolff, M. S. (1998). Estrogenic and antiprogestagenic activities of pyrethroid insecticides. Biochem. Biophys. Res. Commun. 251, 855–859. Gellert, R. J. (1978). Kepone, mirex, dieldrin, and aldrin: estrogenic activity and the induction of persistent vaginal estrus and anovulation in rats following neonatal treatment. Environ. Res. 16, 131–138. Go, V., Garey, J., Wolff, M. S., and Pogo, B. G. (1999). Estrogenic potential of certain pyrethroid compounds in the MCF-7 human breast carcinoma cell line. Environ. Health Perspect. 107, 173–177. Gray, L. E. Jr., and Ostby, J. S. (1995). In utero 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) alters reproductive morphology and function in female rat offspring. Toxicol. Appl. Pharmacol. 133, 285–294. Gray, L. E. Jr., Ostby, J. S., Ferrell, J. M., Sigmon, E. R., and Goldman, J. M. (1988). Methoxychlor induces estrogen-like alterations of behavior and the reproductive tract in the female rat and hamster: effects on sex behavior, running wheel activity, and uterine morphology. Toxicol. Appl. Pharmacol. 96, 525–540. Gray, L. E. Jr., Ostby, J., Ferrell, J., Rehnberg, G., Linder, R., and Cooper, R. (1989). A dose–response analysis of methoxychlor-induced alterations of reproductive development and function in the rat. Fundam. Appl. Toxicol. 12, 92–108. Gray, L. E. Jr., Ostby, J. S., and Kelce, W. R. (1994). Developmental effects of an environmental antiandrogen: the fungicide vinclozolin alters sex differentiation of the male rat. Toxicol. Appl. Pharmacol. 129, 46–52. Gray, L. E. Jr., Wolf, C., Lambright, C., Mann, P., Price, M., and Cooper, R. L. (1999a). Administration of potentially antiandrogenic pesticides (procymidone, linuron, iprodione, chlozolinate, p,p-DDE, and ketoconazole) and toxic substances (dibutyl- and diethylhexyl phthalate, PCB 169, and ethane dimethane sulphonate) during sexual differentiation produces diverse profiles of reproductive malformations in the male rat. Toxicol. Ind. Health 15, 94–118. Gray, L. E. Jr., Ostby, J., Monosson, E., and Kelce, W. R. (1999b). Environmental antiandrogens: low doses of the fungicide vinclozolin alter sexual differentiation of the male rat. Toxicol. Ind. Health 15, 48–64. Griffin, J. E., and Ojeda, S. R. (1988). “Textbook of Endocrine Physiology.” Oxford University Press, New York. Guillette, L. J. Jr., Gross, T. S., Masson, G. R., Matter, J. M., Percival, H. F., and Woodward, A. R. (1994). Developmental abnormalities of the gonad and abnormal sex hormone concentrations in juvenile alligators from contaminated and control lakes in Florida. Environ. Health Perspect. 102, 680–688. Guillette, L. J. Jr., Edwards, T. M., and Moore, B. C. (2007). Alligators, contaminants and steroid hormones. Environ. Sci. 14, 331–347. Guzelian, P. S. (1982). Comparative toxicology of chlordecone (Kepone) in humans and experimental animals. Annu. Rev. Pharmacol. Toxicol. 22, 89–113. Haddow, J. E., Palomaki, G. E., Allan, W. C., Williams, J. R., Knight, G. J., and Gagnon, J. (1999). Maternal thyroid deficiency during pregnancy and subsequent neuropsychological development of the child. N. Engl. J. Med. 341, 549–555. Hadley, M. E. (1996). “Endocrinology.” Prentice Hall, Upper Saddle River, NJ. Hard, G. C. (1998). Recent developments in the investigation of thyroid regulation and thyroid carcinogenesis. Environ. Health Perspect. 106, 427–436. Hayes, T. B., Stuart, A. A., Mendoza, M., Collins, A., Noriega, N., and Vonk, A. (2006). Characterization of atrazine-induced gonadal malformations in African clawed frogs (Xenopus laevis) and comparisons

Chapter | 18  Pesticides as Endocrine-Disrupting Chemicals

with effects of an androgen antagonist (cyproterone acetate) and exogenous estrogen (17beta-estradiol): support for the demasculinization/feminization hypothesis. Environ. Health Perspect. 114(Suppl. 1), 134–141. Herbst, A. L. (1972). Stilbestrol and vaginal cancer in young women. CA Cancer J. Clin. 22, 292–295. Hill, R. N., Crisp, T. M., Hurley, P. M., Rosenthal, S. L., and Singh, D. V. (1998). Risk assessment of thyroid follicular cell tumors. Environ. Health Perspect. 106, 447–457. Hosokawa, S., Murakami, M., Ineyama, M., Yamada, T., Yoshitake, A., and Yamada, H. (1993). The affinity of procymidone to androgen receptor in rats and mice. J. Toxicol. Sci. 18, 83–93. Hotchkiss, A. K., Parks-Saldutti, L. G., Ostby, J. S., Lambright, C., Furr, J., and Vandenbergh, J. G. (2004). A mixture of the “antiandrogens” linuron and butyl benzyl phthalate alters sexual differentiation of the male rat in a cumulative fashion. Biol. Reprod. 71, 1852–1861. Hurley, P. M. (1998). Mode of carcinogenic action of pesticides inducing thyroid follicular cell tumors in rodents. Environ. Health Perspect. 106, 437–445. Janer, G., Slob, W., Hakkert, B. C., Vermeire, T., and Piersma, A. H. (2008). A retrospective analysis of developmental toxicity studies in rat and rabbit: what is the added value of the rabbit as an additional test species? Regul. Toxicol. Pharmacol. 50, 206–217. Jewell, W. T., and Miller, M. G. (1998). Identification of a carboxylesterase as the major protein bound by molinate. Toxicol. Appl. Pharmacol. 149, 226–234. Jewell, W. T., Hess, R. A., and Miller, M. G. (1998). Testicular toxicity of molinate in the rat: metabolic activation via sulfoxidation. Toxicol. Appl. Pharmacol. 149, 159–166. Joshi, S. C., Jain, G. C., and Lata, M. (1994). Effects of ketoconazole (an imidazole antifugal agent) on the fertility and reproductive function of male mice. Acta Eur. Fertil. 25, 55–58. Kavlock, R. J., and Ankley, G. T. (1996). A perspective on the risk assessment process for endocrine-disruptive effects on wildlife and human health. Risk Anal. 16, 731–739. Kavlock, R. J., Daston, G. P., DeRosa, C., Fenner-Crisp, P., Gray, L. E., and Kaattari, S. (1996). Research needs for the risk assessment of health and environmental effects of endocrine disruptors: a report of the U.S. EPA-sponsored workshop. Environ. Health Perspect. 104(Suppl. 4), 715–740. Kelce, W. R., Monosson, E., Gamcsik, M. P., Laws, S. C., and Gray, L. E. Jr. (1994). Environmental hormone disruptors: evidence that vinclozolin developmental toxicity is mediated by antiandrogenic metabolites. Toxicol. Appl. Pharmacol. 126, 276–285. Khera, K. S., Whalen, C., and Trivett, G. (1978). Teratogenicity studies on linuron, malathion, and methoxychlor in rats. Toxicol. Appl. Pharmacol. 45, 435–444. Kim, I. Y., Shin, J. H., Kim, H. S., Lee, S. J., Kang, I. H., and Kim, T. S. (2004a). Assessing estrogenic activity of pyrethroid insecticides using in vitro combination assays. J. Reprod. Dev. 50, 245–255. Kim, Y. S., Katase, T., Sekine, S., Inoue, T., Makino, M., and Uchiyama, T. (2004b). Variation in estrogenic activity among fractions of a commercial nonylphenol by high performance liquid chromatography. Chemosphere 54, 1127–1134. Kim, S. S., Lee, R. D., Lim, K. J., Kwack, S. J., Rhee, G. S., and Seok, J. H. (2005). Potential estrogenic and antiandrogenic effects of permethrin in rats. J. Reprod. Dev. 51, 201–210. Kloas, W., Lutz, I., Urbatzka, R., Springer, T., Krueger, H., and Wolf, J. (2009a). Does atrazine affect larval development and sexual

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Middleton, M. C., Milne, C. M., Moreland, S., and Hasmall, R. L. (1986). Ovulation in rats is delayed by a substituted triazole. Toxicol. Appl. Pharmacol. 83, 230–239. Milne, C. M., Hasmall, R. L., Russell, A., Watson, S. C., Vaughan, Z., and Middleton, M. C. (1987). Reduced estradiol production by a substituted triazole results in delayed ovulation in rats. Toxicol. Appl. Pharmacol. 90, 427–435. Minor, J. L., Knapp, H. F., Stuart, B. O., Killinger, J. M., Zwicker, G. M., and Freudenthal, R. I. (1984). Evaluation of male rat fertility following inhalation exposure to ordram. Toxicologist 4, 80. Monosson, E., Kelce, W. R., Lambright, C., Ostby, J., and Gray, L. E. Jr. (1999). Peripubertal exposure to the antiandrogenic fungicide, vinclozolin, delays puberty, inhibits the development of androgen-dependent tissues, and alters androgen receptor function in the male rat. Toxicol. Ind. Health 15, 65–79. Nelson, J. A., Struck, R. F., and James, R. (1978). Estrogenic activities of chlorinated hydrocarbons. J. Toxicol. Environ. Health 4, 325–339. Noriega, N. C., Ostby, J., Lambright, C., Wilson, V. S., and Gray, L. E. Jr. (2005). Late gestational exposure to the fungicide prochloraz delays the onset of parturition and causes reproductive malformations in male but not female rat offspring. Biol. Reprod. 72, 1324–1335. Ostby, J., Kelce, W. R., Lambright, C., Wolf, C. J., Mann, P., and Gray, L. E. Jr. (1999). The fungicide procymidone alters sexual differentiation in the male rat by acting as an androgen-receptor antagonist in vivo and in vitro. Toxicol. Ind. Health 15, 80–93. Pine, M. D., Hiney, J. K., Lee, B., and Dees, W. L. (2008). The pyrethroid pesticide esfenvalerate suppresses the afternoon rise of luteinizing hormone and delays puberty in female rats. Environ. Health Perspect. 116, 1243–1247. Pont, A., Williams, P. L., Loose, D. S., Feldman, D., Reitz, R. E., and Bochra, C. (1982a). Ketoconazole blocks adrenal steroid synthesis. Ann. Intern. Med. 97, 370–372. Pont, A., Williams, P. L., Azhar, S., Reitz, R. E., Bochra, C., and Smith, E. R. (1982b). Ketoconazole blocks testosterone synthesis. Arch. Intern. Med 142, 2137–2140. Rawlings, N. C., Cook, S. J., and Waldbillig, D. (1998). Effects of the pesticides carbofuran, chlorpyrifos, dimethoate, lindane, triallate, trifluralin, 2,4-D, and pentachlorophenol on the metabolic endocrine and reproductive endocrine system in ewes. J. Toxicol. Environ. Health A 54, 21–36. Reaves, M. E., Cooper, R., Dellarco, V., Dix, D., Martin, M., Mendez, E., and Stoker, T. E. (2008). Retrospective analysis of multigeneration toxicity studies: what is the impact of the second generation in hazard assessment for pesticides? In “The Toxicologist,” pp. 97–102. Baltimore, MD. Robison, L. L. (2009). Treatment-associated subsequent neoplasms among long-term survivors of childhood cancer: the experience of the Childhood Cancer Survivor Study. Pediatr. Radiol. 39(Suppl. 1), S32–S37. Ronis, M. J., and Badger, T. M. (1995). Toxic interactions between fungicides that inhibit ergosterol biosynthesis and phosphorothioate insecticides in the male rat and bobwhite quail (Colinus virginianus). Toxicol. Appl. Pharmacol. 130, 221–228. Ronis, M. J., Ingelman-Sundberg, M., and Badger, T. M. (1994). Induction, suppression and inhibition of multiple hepatic cytochrome P450 isozymes in the male rat and bobwhite quail (Colinus virginianus) by ergosterol biosynthesis inhibiting fungicides (EBIFs). Biochem. Pharmacol. 48, 1953–1965. Shin, J. H., Moon, H. J., Kim, T. S., Kang, I. H., Ki, H. Y., and Choi, K. S. (2006). Repeated 28-day oral toxicity study of vinclozolin in rats

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based on the draft protocol for the “Enhanced OECD Test Guideline No. 407” to detect endocrine effects. Arch. Toxicol. 80, 547–554. Simic, B., Kniewald, J., and Kniewald, Z. (1994). Effects of atrazine on reproductive performance in the rat. J. Appl. Toxicol. 14, 401–404. Sonino, N. (1987). The use of ketoconazole as an inhibitor of steroid production. N. Engl. J. Med. 317, 812–818. Stevens, J. T., Breckenridge, C. B., Wetzel, L. T., Gillis, J. H., Luempert, L. G. 3rd, and Eldridge, J. C. (1994). Hypothesis for mammary tumorigenesis in Sprague–Dawley rats exposed to certain triazine herbicides. J. Toxicol. Environ. Health 43, 139–153. Stoker, T. E., Goldman, J. M., and Cooper, R. L. (1993). The dithiocarbamate fungicide thiram disrupts the hormonal control of ovulation in the female rat. Reprod. Toxicol. 7, 211–218. Stoker, T. E., Cooper, R. L., Goldman, J. M., and Andrews, J. E. (1996). Characterization of pregnancy outcome following thiram-induced ovulatory delay in the female rat. Neurotoxicol. Teratol. 18, 277–282. Stoker, T. E., Laws, S. C., Guidici, D. L., and Cooper, R. L. (2000). The effect of atrazine on puberty in male Wistar rats: an evaluation in the protocol for the assessment of pubertal development and thyroid function. Toxicol. Sci. 58, 50–59. Stoker, T. E., Jeffay, S. C., Zucker, R. M., Cooper, R. L., and Perreault, S. D. (2003). Abnormal fertilization is responsible for reduced fecundity following thiram-induced ovulatory delay in the rat. Biol. Reprod. 68, 2142–2149. Stoker, T. E., Perreault, S. D., Bremser, K., Marshall, R. S., Murr, A., and Cooper, R. L. (2005). Acute exposure to molinate alters neuroendocrine control of ovulation in the rat. Toxicol. Sci. 84, 38–48. Stoker, T. E., Zorrilla, L. M., Gibson, E. K., Cooper, R. L. (2009). Triclosan exposure modulates estrogen-dependent responses in the rat. Endocrine Society Presented at Washington, DC. Sturtz, N., Deis, R. P., Jahn, G. A., and Duffard, R. (2008). Evangelista de Duffard A.M. Effect of 2,4-dichlorophenoxyacetic acid on rat maternal behavior. Toxicology 247, 73–79. Swartz, W. J., and Corkern, M. (1992). Effects of methoxychlor treatment of pregnant mice on female offspring of the treated and subsequent pregnancies. Reprod. Toxicol. 6, 431–437. Swartz, W. J., Wink, C. S., and Johnson, W. D. (1994). Response of adult murine uterine epithelium to 50% methoxychlor. Reprod. Toxicol. 8, 81–87. Tamura, H., Maness, S. C., Reischmann, K., Dorman, D. C., Gray, L. E., and Gaido, K. W. (2001). Androgen receptor antagonism by the organophosphate insecticide fenitrothion. Toxicol. Sci. 60, 56–62. Tennant, M. K., Hill, D. S., Eldridge, J. C., Wetzel, L. T., Breckenridge, C. B., and Stevens, J. T. (1994a). Chloro-s-triazine antagonism of estrogen action: limited interaction with estrogen receptor binding. J. Toxicol. Environ. Health 43, 197–211. Tennant, M. K., Hill, D. S., Eldridge, J. C., Wetzel, L. T., Breckenridge, C. B., and Stevens, J. T. (1994b). Possible antiestrogenic properties of chloro-s-triazines in rat uterus. J. Toxicol. Environ. Health 43, 183–196. Thakur, A. K., Wetzel, L. T., Voelker, R. W., and Wakefield, A. E. (1998). Results of a two-year oncogenicity study in the Fischer 344 rats with atrazine. In “Triazine Herbicides Risk Assessment” (J. E. M. L. G. Ballentine and D. S. Hackett, Eds.). American Chemical Society, Washington, DC. Trentacoste, S. V., Friedmann, A. S., Youker, R. T., Breckenridge, C. B., and Zirkin, B. R. (2001). Atrazine effects on testosterone levels and androgen-dependent reproductive organs in peripubertal male rats. J. Androl. 22, 142–148. U.S. EPA (1998a). Endocrine Disruptor Screening and Testing Advisory Committee (EDSTAC) Final Report.

Chapter | 18  Pesticides as Endocrine-Disrupting Chemicals

U.S. EPA (1998b). Endocrine Disruptor Screening Program; Proposed Statement of Policy. Federal Register, pp. 71542–71568. U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (2009a). “Safe Drinking Water Act (SWDA). Basic Information 2004.” http://www.epa.gov/safewater/sdwa/basicinformation.html. U.S. EPA (2009b). “Federal Insecticide, Fungicide, and Rodenticides Act (FIFRA), 1997.” http://www.epa.gov/compliance/civil/fifra/fifraenfstatreq.html. U.S. EPA (2009c). “Endocrine Disruptor Screen and Testing Advisory Committee (EDSTAC). Final Report.” http://www.epa.gov/scipoly/ oscpendo/pubs/edspoverview/edstac.html. U.S. EPA (2009d). “Overview of the April 2009 Final List of Chemicals for Initial Tier 1 Screening, Endocrine Disruptor Screening Program (EDSP).” http://www.epa.gov/scipoly/oscpendo/pubs/prioritysetting/ final_listfacts.htm. Van Ravenzwaay, B. (1992). Discussion of prenatal and reproduction toxicity of Reg. No. 83  258 (Vinclozolin). Vinggaard, A. M., Nellemann, C., Dalgaard, M., Jorgensen, E. B., and Andersen, H. R. (2002). Antiandrogenic effects in vitro and in vivo of the fungicide prochloraz. Toxicol. Sci. 69, 344–353. Vinggaard, A. M., Christiansen, S., Laier, P., Poulsen, M. E., Breinholt, V., and Jarfelt, K. (2005). Perinatal exposure to the fungicide prochloraz feminizes the male rat offspring. Toxicol. Sci. 85, 886–897.

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Waller, D. P., Martin, A., Vickery, B. H., and Zaneveld, L. J. (1990). The effect of ketoconazole on fertility of male rats. Contraception 41, 411–417. Walters, L. M., Rourke, A. W., and Eroschenko, V. P. (1993). Purified methoxychlor stimulates the reproductive tract in immature female mice. Reprod. Toxicol. 7, 599–606. Waters, K. M., Safe, S., and Gaido, K. W. (2001). Differential gene expression in response to methoxychlor and estradiol through ERalpha, ERbeta, and AR in reproductive tissues of female mice. Toxicol. Sci. 63, 47–56. Wickramaratne, G. A., Foster, J. R., Ellis, M. K., and Tomenson, J. A. (1998). Molinate: rodent reproductive toxicity and its relevance to humans—a review. Regul. Toxicol. Pharmacol. 27, 112–118. Wilson, V. S., Bobseine, K., and Gray, L. E. (2004). Development and characterization of a cell line that stably expresses an estrogenresponsive luciferase reporter for the detection of estrogen receptor agonist and antagonists. Toxicol. Sci. 81, 69–77. Zachow, R., and Uzumcu, M. (2006). The methoxychlor metabolite, 2,2bis-(p-hydroxyphenyl)-1,1,1-trichloroethane, inhibits steroidogenesis in rat ovarian granulosa cells in vitro. Reprod. Toxicol. 22, 659–665. Zorrilla, L. M., Gibson, E. K., Jeffay, S. C., Crofton, K. M., Setzer, W. R., and Cooper, R. L. (2009). The effects of triclosan on puberty and thyroid hormones in male Wistar rats. Toxicol. Sci. 107, 56–64.

Chapter 19

Volatile Organic Compounds from Pesticide Application and Contribution to Tropospheric Ozone Melanie Marty1, ��Frank �������� Spurlock ���������2 and Terrell Barry2 1

Office of Environmental Health Hazard Assessment, California Environmental Protection Agency, Oakland, California Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California

2

19.1  Introduction Volatile organic compounds (VOC) are organic chemicals that when released into the atmosphere can react with sunlight and nitrogen oxides (NOx) to form tropospheric (ground-level) ozone. Ground-level ozone is one of the six “criteria pollutants” identified in the United States Federal Clean Air Act for which U.S.EPA sets National Ambient Air Quality Standards (NAAQS). Of these six, particulates and ground-level ozone are considered the most widespread health threats (http://www.epa.gov/air/urbanair/). The NAAQS for ozone is 0.075 ppm (8 h average). At high concentrations, tropospheric ozone (O3) causes respiratory problems and is detrimental to a wide variety of plants. The primary ground-level ozone-forming process is (Eq. 1) photolysis of NO2, followed by (Eq. 2) O3 formation from reaction of atomic oxygen with O2, and (Eq. 3) subsequent reaction between O3 and NO to yield NO2 (NRC, 1991).

NO2 hν → O  NO

(1)



O  O 2 → O3

(2)



O3  NO → NO2  O

(3)

Under normal (unpolluted) conditions, O3 concentrations are relatively low because reaction (Eq. 3) prevents any buildup of O3. However, when reactive VOCs are present, they can photolyze to form radicals that react with NO. This slows O3 consumption via (Eq. 3), causing tropospheric O3 concentrations to increase. While the detailed chemistry is far more complex than the few reactions Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

shown here, this simplified description illustrates some important features of tropospheric O3 chemistry. The rate of O3 formation is dependent on the rate of NO2 photolysis, which depends on sunlight intensity. This is one reason why the highest O3 concentrations are observed during California’s May–October “ozone season.” l VOCs do not participate directly in the O3 formation reactions (Eqs. 1 and 2), but instead promote the rate that O3 is formed from the photolysis of NO2. l All VOCs are not created equal with regard to promoting ozone formation. This is because VOCs react at different rates and have differing effects on the radicals that react with NO and promote ozone formation. l The rate of O3 formation depends not only on VOC concentrations, but also on concentrations of nitrogen oxides (NOx). l

Areas that have failed to meet U.S.EPA’s O3 standards are known as nonattainment areas (NAA). California has five O3 NAAs classified as “serious,” “severe,” or “extreme” (http://www.epa.gov/region/air/maps/maps_ top.html). The CAA requires states to develop State Implementation Plans (SIP) that describe how the states will achieve compliance with air quality standards in NAAs. Approved plans include reduction goals and a timeline for achieving those goals. The “pesticide element” of California’s SIP required the California Department of Pesticide Regulation (CDPR) to develop and maintain a pesticide emission inventory for the purpose of tracking progress in achieving VOC reductions. CDPR began developing the United States’ first emission inventory to estimate pesticide VOC emissions in the mid-1990s 571

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572

(http://www.cdpr.ca.gov/docs/emon/vocs/vocproj/vocmenu.htm). The pesticide inventory includes emissions from agricultural and commercial structural pesticide applications. Home-use pesticides are not included in CDPR’s inventory because their emissions are regulated by California’s Air Resources Board. In California’s agriculturally productive San Joaquin Valley NAA and Ventura NAA, pesticides are among the top 10 sources, contributing approximately 5% of emissions in each. The various 2006 O3 season San Joaquin Valley VOC sources are shown in Table 19.1.

19.2  Principal voc contributing pesticides A broad range of pesticide types and formulations are used in California, but two general classes of products contribute the vast majority of pesticidal VOC emissions: fumigants and emulsifiable concentrates (EC). In California’s San Joaquin Valley NAA, fumigants generally contribute 30–40% of total O3 season emissions while ECs account for about 50%. The high contribution of fumigants to emissions is attributable to their relatively widespread use in a broad

Table 19.1  Contributors to 2006 San Joaquin Valley Ozone Season Emissions. Total May–October Estimated Emissions from All Sources were  450 tons Per Day Category

Emissions (tons per day)

Fraction of total (%)

Passenger vehicles

62

13.7

Other (waste disposal/ composting)

57

12.6

Livestock waste (dairy cattle)

40

8.9

Oil and gas production (evaporative losses/flaring)

28

6.2

Consumer products

24

5.3

Pesticides

22

4.9

Heavy duty diesel trucks

20

4.4

Recreational boats

20

4.4

Food and agriculture (crop processing and wineries)

13

2.9

Architectural coatings (paints and thinners)

11

2.4

154

34.1

All other sources

range of crops as a pre-plant soil treatment, high application rates (on the order of 200 kg/ha), and high volatility. Most fumigant products do not contain “inert” ingredients, so fumigant emissions are typically either from the fumigant active ingredients (AI) themselves or their immediate degradates. The principal fumigant AIs contributing to VOC emissions are methyl bromide, 1,3-dichloropropene, chloropicrin, metam-sodium, potassium N-methyl dithiocarbamate (metam-potassium), dazomet, and sodium tetrathiocarbonate. In contrast to fumigants, the composition of ECs are usually dominated by hydrocarbon solvents and carriers. Many of these non-AI ingredients are sufficiently volatile to contribute to emissions, and they are often present in amounts greater than the actual AIs. In a few cases, the active ingredient may contribute substantially to emissions [e.g. the aquatic herbicide acrolein, molinate (no longer registered), EPTC, and metaldehyde]. However, it is clear that a large majority of EC emissions are attributable to formulation components such as solvents.

19.3  CDPR initial method to obtain a screening assessment of voc mass released from each california pesticide—early 1990s Thermogravimetric analysis (TGA) was selected in 1991 as a method to determine the VOC emissions potential of pesticide applications because it is versatile, widely used, requires a small sample mass, and has a short turnaround time. The CDPR TGA method for pesticides was developed using a protocol (Pino and Barry, 1992) based upon ASTM D-3960 (ASTM, 1989).

19.3.1  Method Development A TGA pilot study was conducted on three pesticide formulations: (1) emulsifiable concentrate; (2) liquid; and (3) dry flowable. The statistical technique of Response Surface Methodology (RSM) was used to optimize analytical conditions. In the optimal method, a 10-mg sample is heated to 115°C (ramped at 5°C/min) and held until mass loss is stabilized at 0.05% for 5 min (Pino et al., 1996). A TGA scope investigation study using this protocol was conducted on 18 pesticide products from 13 formulation cat­egories. Results confirmed that the parameters optimized in the TGA pilot study were suitable for analysis of a broad range of pesticide products.

19.3.2  Method Validation CDPR planned and executed a two-phase collaborative study to validate the method and estimate the method

Chapter | 19  Volatile Organic Compounds from Pesticide Application and Contribution to Tropospheric Ozone

repeatability and reproducibility (Pino and Barry, 1992). The two phases consisted of a performance trial (PT) and an interlaboratory trial (ILT). The objectives of the PT were: (1) detect problems with the written description of the method; (2) assess the performance of the prospective interlaboratory trial laboratories; and (3) obtain a prelimin­ ary estimate of the method repeatability and reproducibility. For the PT, 17 laboratories analyzed Youden Pairs (Wernimont, 1985; Youden and Steiner, 1975) of two liquid pesticide formulations (a liquid solution and an EC). The Youden Pair is a set of two samples of the same pesticide formulation where one sample is diluted with a small amount of solvent. The advantage of Youden Pairs is that it is impossible for laboratories to censor their results because there are no two samples exactly alike. Preliminary estimates of repeatability were 3.08% and 0.73%, and of reproducibility were 8.85% and 1.21% absolute for the two pesticide formulations, respectively (Pino et al., 1996). The second phase, the ITL, fully characterized the method’s repeatability and reproducibility. Seventeen la­boratories analyzed three standards (low, medium, and high volatility) and Youden Pairs of five pesticide formulations (two granular, a suspension, a flowable, and a liquid solution). Results for the standards indicate that the TGA method is unbiased over the entire range of potential percent mass loss. Mean mass loss for the standards differed by less than 1% from target values. The pesticide formulation repeatability estimates varied from 0.60% absolute for the high volatility liquid solution to 6.28% absolute for the coarse granular formulation. The pesticide formulation reproducibility estimates varied from 0.75% absolute for the high volatility liquid solution to 6.77% absolute for the coarse granular formulation. Heterogeneous formulations showed larger variability than homogeneous formulations. In addition, homogeneous formulations that are less volatile tend to be more variable than highly volatile formulations (Pino et al., 1996).

19.3.3  Corrections to TGA Results Water and volatiles exempted by U.S.EPA are included in the TGA volatilization mass loss measurements. Exempt chemicals have been determined by U.S.EPA to display negligible photochemical reactivity and are listed in 40 CFR51.000(s) (http://epa.gov/ttn/naaqs/ozone/ozonetech/def_voc.htm). Thus, the TGA test results were corrected to obtain the final VOC emission potential estimate using a simple subtraction method for both water and exempt volatiles.

19.3.4  Default Volatility Values In the mid-1990s CDPR requested that pesticide registrants submit thermogravimetric analytical (TGA) data for all

573

products that were then registered for agricultural or commercial structural uses. These data are used to determine product emission potentials (EP). However, the data callin was voluntary, and registrants submitted TGA data for only a fraction of then-registered products. CDPR developed several alternate methods to estimate product EPs for those products without TGA data, including water subtraction, inorganic subtraction, and the assignment of default EPs. The “subtraction” methods assumed that all product components except water or inorganic materials contributed to VOC emissions. The default EPs were defined for each formulation category (e.g. emulsifiable concentrate, flowable concentrate, dusts, and powders). Default EPs were initially defined as the highest approved TGA EP in a formulation category, partly to encourage pesticide registrants to provide TGA data for their products. However, the assignment of the arbitrary high default EP values was ineffective for encouraging submission of TGA data and resulted in substantial error in the inventory (Spurlock, 2002a). The high default EPs resulted in upward bias of early inventory estimates, and did not allow meaningful analysis of year-to-year trends in emissions. These problems were subsequently resolved by re-defining the default EPs as the median TGA-based EP in formulation class (Spurlock, 2002b).

19.3.5  VOC Inventory Calculations In California, all agricultural and commercial pesticide applications must be reported. The county agricultural commissioners and CDPR compile the reports into a database, the Pesticide Use Report (PUR). CDPR then uses the pounds of pesticide product applied that is shown in the PUR to calculate the VOC emissions for each pesticide product as: VOC emission (pounds)  pounds product applied  EP The total VOC emission for a particular nonattainment area is the sum of the VOC emissions for all products in that attainment area. The first pesticide VOC inventory was calculated for the years 1990–1995 (Oshima, 1998). Due to a limited response to the data call-in, a significant proportion of the products had EPs calculated by “water subtraction” or were assigned the default EP value for the formulation class. The 1990 base year had the maximum number of products that used the “maximum-value default EP.” As a result, the inventory base year of 1990 showed high VOC emissions relative to the 1990 pesticide use. This was later corrected by revisions and improvements to the early program. CDPR has now obtained TGA data for many more products. Default emission potentials now account for only 5–10% of total emissions in current years.

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19.4  Revisions and improvements to the early program After adoption of the TGA method for measuring pesticide product emission potentials, several refinements were made to the pesticide VOC program. Most of these changes were in two key areas: development of “special default” EPs for certain product classes that accounted for their particular chemistry, and adoption of fumigant field adjustments to account for the effect of application method on emissions.

19.4.1  Special Default Emission Potential Metam-sodium and potassium N-methyl dithiocarbamate are unusual in the sense that both are nonvolatile salts. Their fumigant action is due to their hydrolysis product, methyl isothiocyanate (MITC), which is rapidly formed post application. Consequently the TGA method is an inappropriate EP estimation method for these products, and CDPR defined emission potentials for these two products assuming stoichiometric conversion of the applied AI to MITC as follows: metam sodium EP  product mass fraction metam sodium  0.566. Potassium N-methyl dithiocarbamate EP  product mass fraction potassium N-methyl dithiocarbamate  0.503. l Dazomet is formulated as a granular product, and is also a nonvolatile MITC-generating fumigant. Research on the environmental fate of dazomet indicates that a variety of mixed volatile organics can result from dazomet breakdown (Subramanian et al., 1996). Consequently dazomet product EP  mass fraction dazomet in product. l Sodium tetrathiocarbonate is a fumigant that degrades rapidly post-application to the actual pesticidal agent carbon disulfide. The EP for sodium tetrathiocarbonate products are expressed on a carbon disulfide equivalent basis assuming stoichiometric conversion to carbon disulfide: EP  product mass fraction sodium tetrathiocarbonate  0.409. l Several inorganic pesticides are widely used in California for commercial structural use, post-harvest commodity treatment or other miscellaneous uses (e.g. Cl2, CO2). Products containing these chemicals do not contain organic carbon so they were assigned EP  0. l Sodium chlorate defoliant products are liquid formulations, and confidential statements-of-formula (CSF) submitted show the products contain only water, alumin­ osilicates (clay), inorganic salts, and/or urea fire retardant. Urea is highly polar, nonvolatile, hydrophilic, and has a very low Henry’s law constant, and preferentially partitions into water or moist soil where it degrades rapidly to ammonia and carbon dioxide. Consequently, CDPR assigned all sodium chlorate products EP  0. l

Large quantities of elemental sulfur are applied as a fungicide in a variety of crops, often as fine powders mixed with clay. Their TGA-based EPs are generally quite low, typically around 1–2%. However, extremely large amounts are applied so they were initially considered major contributors, yielding emissions in the range of 1–2 tons per day. CDPR-funded research identifed the volatile components of six representative dry sulfur products during TGA analysis as water vapor. Based on those data, sulfur dust and powder products are now assigned EP  0 providing no other organic chemicals are present based on product CSFs.

l

19.5  Fumigant adjustments for field application method In 2007, CDPR issued a memorandum outlining methods used to develop emission adjustment factors (AF) for fumigants (Barry et al., 2007). The application method AF is the proportion of the fumigant mass applied that is emitted to the air following application. AFs are specific for each fumigant/application method combination, are based upon measured data, and yield refined estimates of fumigant VOC emissions. Prior to 2007 the EP for all fumigants was 100. However, CDPR has several dozen field studies for various fumigant/application method combinations that show VOC emissions ranging from 9% to 100%. Where data were not available, emissions were estimated with surrogate data. Based on various data, the fraction of fumigant applications by each application method were estimated for the 1990 base year and for each year since 2004. Emissions are then calculated as follows: VOC emission (pounds)  pounds product applied EP  application method AF

19.5.1  Methyl Bromide (MB) The CDPR MB database includes 30 field studies utilizing current application methods. CDPR’s analysis of these data (Barry, 1999; Barry et al., 2007) resulted in three distinct groups of MB application methods and AFs were calculated for each group: (1) shank injection into pre-formed beds followed by covering with tarpaulin (AF  1.0); (2) shank injection into flat fields (broadcast) followed by covering with tarpaulin (AF  0.74); and (3) shank injection into flat fields (broadcast) without a tarpaulin (AF  0.48).

19.5.2  1,3-Dichloropropene (1,3-D) The CDPR 1,3-D database includes six studies. Four studies used shank injection at varying depth, with and without tarpaulin covering. Two studies used the drip application

Chapter | 19  Volatile Organic Compounds from Pesticide Application and Contribution to Tropospheric Ozone

method. Analysis of the 1,3-D data resulted in four distinct application method groups: (1) shallow injection into flat fields (broadcast) (AF  0.65); (2) shallow injection with three irrigations (AF  0.44); (3) deep injection into flat fields (broadcast) (AF  0.26); and (4) drip irrigation (AF  0.29). Unlike MB, 1,3-D data showed no reduction in emissions with tarpaulin covering but showed an apparent reduction in emissions with shank injection depth (Johnson, 2006).

19.5.3  Chloropicrin (CP) Chloropicrin is used most often as a mixture with MB or 1,3-D. However there are some applications of CP as the sole fumigant. The application methods for CP are the same methods as for MB and 1,3-D, but the AF values are unique to CP (Barry et al., 2007). The CP emissions are most affected by the type of tarpaulin (high versus low permeability) or the lack of a tarpaulin following application. Since MB bed applications have high emissions (Barry, 1999), the CP bed method was combined with the no /high permeability tarpaulin methods. Drip application method shows the lowest emission rate. The CP application method groups are: (1) shallow or deep shank injection bed or broadcast covered with high permeability tarpaulin (AF  0.64); (2) shallow or deep shank injection, broadcast covered with low permeability tarpaulin (AF  0.44); (3) deep injection, no tarpaulin, three irrigations (AF  0.43); and (4) drip irrigation covered by low or high permeability tarpaulin (AF  0.12).

19.5.4  Metam Sodium Metam sodium (MS) and metam potassium (MP) fumigant action and VOC emissions are due to the hydrolysis product methyl isothiocyanate (MITC), which is generated when sufficient water is applied following application. The EPs for these two products are expressed on an MITC equivalent basis (Spurlock, 2005). The AFs are also on an MITC equivalent basis. The MS database consists of studies conducted by the Metam Sodium Task Force (MSTF) under their 1997–2001 Field Program, and CDPR (Barry et al., 2007). The data indicate that post-application irrigation reduces emissions substantially, and drip and rototiller application methods have low emissions. The MS application method groupings were: (1) shallow shank injection or sprinkler with or without tarpaulin, and flood (AF  0.77); (2) shallow shank injection or sprinkler with two post-post application irrigations (AF  0.28); (3) shallow shank injection or sprinkler with three post-post application irrigations (AF  0.21); (4) rototill or shank injection followed by soil-capping (AF  0.14); and (5) drip irrigation with or without tarpaulin (AF  0.09).

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19.5.5  Dazomet The fumigant action of dazomet is also due to the generation of MITC, but dazomet degradation also results in a variety of other mixed organics (Subramanian et al., 1996). Thus, the EP for dazomet is expressed on an AI basis instead of an MITC equivalent basis. CDPR has three dazomet studies, and the emissions differ by an order of magnitude. Due to uncertainty in the true emissions, CDPR is using the average fraction of MITC emitted as the interim AF value for all application methods (AF  0.17) (Barry et al., 2007).

19.5.6  Sodium Tetrathiocarbonate Since the fumigant action and VOC emission of sodium tetrathiocarbonate are due to the hydrolysis product carbon disulfide, the EP for sodium tetrathiocarbonate is expressed on a carbon disulfide equivalent basis (Spurlock, 2006). The mini-sprinkler application method represents worst case (Haskell, 1995). The AF of 0.10 is used for mini-sprinkler, drip, and flood applications of sodium tetrathiocarbonate (Barry et al., 2007).

19.6  Future issues On January 18, 2008 CDPR issued a notice of reevalu­ ation for certain pesticides products including MB, 1,3D, CP, MS, MP, dazomet, and sodium tetrathiocarbonate. Registrants of products containing those fumigants are required to submit flux monitoring studies, either direct flux or back calculation, for field fumigation application methods (Johnson et al., 1999). The results from these studies will be used to further refine the original AFs in Barry et al., (2007).

19.6.1  Adjusting Nonfumigants for Application Method Effects on Emissions Nonfumigant pesticides may be applied foliarly, to water, to the soil surface or incorporated into the soil. While TGA provides a consistent, well-defined measure of a product’s volatility, under field conditions the effects of sorption to soil and/or plants and degradation are expected to decrease actual emissions in some cases relative to the laboratorybased TGA method. It would be desirable to account for environmental conditions and the effect of application method for nonfumigants, similar to the adjustments made for fumigants. Very little actual field data on ultimate volatilization of nonfumigant products exists. In theory, modeling is one avenue that could be used to estimate adjustment factors for nonfumigants, but the lack of field data for validating models is problematic.

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One reason why there is so little data on ultimate volatilization of nonfumigant products is the difficulty in conducting such studies. Nonfumigant products such as emulsifiable concentrates often contain mixtures of solvents and dozens of potentially volatile components. Consequently, post-application monitoring and chemical analysis of many of the components over time would be required to characterize fate and ultimate volatilization of the product under field conditions. A second problem is that application rates for nonfumigant products are generally much lower than fumigants, so analytical detection limits create problems in trying to measure slow rates of volatilization. A third problem is the length of time required to determine ultimate volatilization of nonfumigant products. Lower volatility solvents may continue to volatilize in the field over time frames of weeks up to a month or more. Finally, as required by the CAA, California’s SIP sets pesticide VOC reduction goals relative to base year emissions in 1990. Thus, any adjustments to nonfumigant emissions to account for field conditions would necessarily have to be developed for both the base year as well as for current years. California has a comprehensive use reporting system, but pesticide use reports contain almost no information on method of application. Given that CDPR’s 1990–2007 VOC inventory includes more than 10,000 nonfumigant products containing several hundred active ingredients, it would be very difficult to retroactively classify application methods for the many thousands of crop/ product combinations. Thus, adjustments to nonfumigant emissions that account for potential effects of field conditions are unlikely to be developed any time soon.

SAPRC-07 (Carter, 2009), where the model is evaluated against concentration time series data measured in envir­ onmental chambers under well-controlled conditions of temperature, light intensity, NOx, and VOC composition (Carter and Malkina, 2007). Table 19.2 illustrates the variability of MIRs for various pesticide product components. Interestingly methyl bromide has a MIR lower than ethane. U.S.EPA has used ethane’s reactivity as a cutoff for classifying a VOC’s reactivity as negligible and exempted from regulation (Dimitriades, 1999). No exemption decision has been made to date for methyl bromide. While methyl bromide is currently being phased out over concerns about depeletion of the stratospheric O3 layer, the methyl bromide substitutes 1,3-dichloropropene and chloropicrin have much greater tendencies to form tropospheric O3.

Table 19.2  Maximum Incremental Reactivities (g O3/g VOC) of Selected Pesticide Components as Reported by Carter and Malkina (2007) Product component or degradate

VP (ppm)

MIR

Methyl bromide

1000

0.03

MITC (methyl isothiocyanate)

1000

0.35

1,3-Dichloropropene

1000

4.64

Chloropicrin

1000

2.18

Molinate

7.4

1.68

Kerosene

–

1.71

19.6.2  Speciation/Reactivity

Methylisobutyl ketone

1000

4.28

A major scientific and regulatory issue with California’s pesticide VOC inventory is that it is mass-based. VOCs are highly variable in their ability to promote ozone formation. Consequently a mass-based inventory may result in regulatory inefficiency. Regulatory focus may be diverted from the most problematic pesticide product constituents to those which actually have minor or no O3 formation potential. Reformulation of high VOC emulsifiable concentrates to lower VOC products could yield a lower total mass of VOC emissions but actually increase O3 formation. Several reactivity scales have been developed to describe the relative ability of different VOCs to form O3 (Carter, 1994; Derwent, 2004). One of the most common is the maximum incremental reactivity (MIR) scale, expressed in units of (gm O3 formed)/(gm VOC added). The MIR describes the marginal change in O3 resulting from addition of that VOC to a well-defined “base” mixture consisting of NOx and other VOCs representative of O3 forming conditions (Carter, 1994). MIRs are typically calculated using a chemical mechanism model such as

Acrolein

1000

7.55

Glycerine

0.22

3.26

Propylene glycol

170

2.74

Thiobencarb

0.03

0.72

N-Methyl pyrrolidinone

454

2.55

Ethyl di-n-propyl-thiolcarbamate (EPTC)

32

1.82

Pebulate

116

1.84

1000

0.28

1000

3.71

1000

0.31

Carbon disulfide a

Base ROG mixture Ethane

b

MIR  maximum incremental reactivities; VP  vapor pressure. a Mixture used to represent reactive organic gas (ROG) emissions from all sources for the purpose of calculating atmospheric ozone impacts (Carter, 1994). b Ethane has been used by the U.S.EPA to define the borderline between reactive and negligible reactivity for VOC exemption purposes.

Chapter | 19  Volatile Organic Compounds from Pesticide Application and Contribution to Tropospheric Ozone

Under high NOx and low VOC conditions, O3 formation is limited by VOC concentrations. Conversely, with low NOx and high VOC concentrations, O3 formation is NOx limited. One criticism of the MIR scale as applied to agricultural regions is that MIRs are determined for base case scenarios that reflect urban high NOx conditions. By contrast, rural agricultural areas have much lower NOx, attributable in part to less vehicle traffic. Because O3 formation under rural conditions may often be NOx limited, MIRs determined under high NOx base case scenarios may not accurately reflect VOC reactivities. However, the widespread use and availability of MIR data for VOCs favors the use of that scale until reactivity data becomes widely available for low NOx scenarios. CDPR is just beginning to consider how reactivity may be incorporated into the pesticide VOC inventory. The first issue is that both the actual product component chemicals and the composition of a product’s VOC emissions (speciation) need to be identified. For fumigants, speciation is relatively easy because the product consists almost entirely of active ingredient. For certain liquid products, the fraction of volatile components (e.g. solvents) in a product is similar to the TGA value. For these, the composition of the solvent yields composition of the VOC emissions. In theory, product compositions are given by CSFs submitted at the time of registration. However, CSFs often list components and their alternates by commercial names. Many solvents used in formulation are processed blends, distillation cuts, or have proprietary compositions. Ultimately the goal would be speciation of product emissions, and subsequent application of component MIRs to estimate a weighted MIR for product emissions. That reactivity data can then be used to focus regulatory efforts on products that have the greatest contribution to O3 formation.

19.6.3  Ozone Toxicity The primary target organ for O3 health effects is the re­spiratory tract including associated immune system cells. While systemic effects of O3 exposure have been reported, this section focuses on the lung. There is a large literature on effects of O3 exposure in both animals and humans; this section should be viewed as a brief synopsis of the available information. Both the U.S. Environmental Protection Agency criteria document for the National Ambient Air Quality Standards (U.S.EPA, 2006) and the California Environmental Protection Agency (Cal/EPA) Review of the California Ambient Air Quality Standard for Ozone (Cal/ EPA, 2005) provide exhaustive reviews of the literature.

19.6.4  Mechanisms of Toxicity The toxicity of O3 is related to its oxidant properties. O3 induces lipid peroxidation, a free radical chain reaction

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involving membrane lipids. While lung tissue and lining fluid have stores of antioxidant, such as reduced glutathione (GSH), these stores can be depleted resulting in cell damage and death. Inflammation, which underlies the respiratory toxicity of ozone, involves a cascade of biochemical events following injury, including the release of pro-inflammatory cytokines by epithelial cells and macrophages, and influx of inflammatory cells into the alveolar and bronchiolar interstitium (Cal/EPA, 2005). Enhancement of tissue injury also occurs from release of proteolytic enzymes and reactive oxygen species (ROS) from inflammatory cells, resulting in additional oxidative damage to membrane lipids and proteins. O3 also reacts with lipids in the epithelial lining fluid to produce lipid ozonation products (LOPs) (Ciencewicki et al., 2008). LOPs contribute to oxidative stress, activate phospholipases A2, C, and D, and induce release of inflammatory cytokines including IL-6, IL-8, and prostaglandin-E2. O3 and ROS produced during cell injury also oxidize proteins in the lung fluid and epithelial cell membranes. Surfactant proteins can be destroyed by ROS decreasing the protective action of these proteins against inflammation and further oxidation, and compromising alveolar patency. O3 may produce changes in lung function by stimulating C-fibers that are vagal afferents in the airway (Cal/EPA, 2005). Increased substance P, a neuropeptide released by bronchial C fibers, is observed in bronchoalveolar lavage of ozone-exposed subjects. O3-induced C fiber stimulation inhibits inspiration, resulting in the characteristic rapid shallow breathing seen in animals and humans following ozone exposure. Bronchoconstriction may also involve stimulation of the parasympathetic nervous system controlling airway caliber.

19.6.5  Animal Studies 19.6.5.1  Increased Epithelial Permeability Movement of particle and macromolecules across the airway epithelium is normally restricted by tight junctions between epithelial cells. O3 exposure disrupts this barrier resulting in increased permeability of serum proteins and fluid into the air spaces and transport of exogenous material from the air spaces into the interstitium and blood. Increased radiolabelled albumin was observed in bronchoalveolar lavage fluid (BALF) after continuous exposure to 0.2 ppm O3 for 2 days or 0.4 ppm for 6 h (Guth et al., 1986). Transient increases in tracheal and bronchoalveolar permeability were observed following 2–3 h exposures of rats and guinea pigs to 0.8–1.0 ppm O3 (Bhalla and Crocker, 1987; Miller et al.,1986). Exposure in vitro of rat alveolar epithelial monolayers to a range of O3 concentrations (0.1–1.0 ppm) resulted in a dose-dependent decrease in barrier function due to epithelial cell degradation (Cheek et al., 1994).

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19.6.5.2  Lung Structure and Cellularity Changes A number of investigators have observed statistically significant increases in alveolar macrophages (AM), protein, albumin, total cell counts, and polymorphonuclear monocytes (PMNs) numbers in BALF and bronchiolar epithelium of rodents, dogs, and monkeys at various time points following acute exposures to 0.2 ppm O3 (Freed et al., 1999; Kleinman et al., 1999; Oosting et al., 1991; Plopper et al., 1998). Both monkeys and ferrets appear to be more sensitive to O3-induced inflammatory cell infiltration and airway epithelial injury than rodents (Sterner-Kock et al., 2000). Adaptation to some of the acute effects of O3 exposure, such as rapid shallow breathing, has been observed in a number of animal studies and humans (Folinsbee et al., 1980), but biochemical and morphological changes continue to progress with repeated exposure (Tepper et al., 1989). Tepper noted attenuation of lung function change but progression of epithelial damage with repeat 2.5-h exposures to 0.5 ppm O3 on consecutive days. While the acute inflammatory response measured as increased PMNs and protein in BALF appeared to resolve a few days into a chronic exposure in rats to 0.4 ppm (van Bree et al., 2001), numbers of AM continued to increase during the course of chronic exposure. Many studies that evaluated lung pathology during and after repeated or chronic O3 exposures in the rat report excess collagen synthesis and fibrosis in the centriacinar region (Boorman et al., 1995; Last et al., 1979, 1993). Chronic exposure inhibits tissue repair, possibly by depressed cell proliferation, and structural remodeling ensues. The most pronounced lesions occur in the centriacinar region which receives the largest O3 dose (e.g. terminal and respiratory bronchioles in primates) (Plopper et al., 1998). Ciliated respiratory epithelium in the tracheobronchiolar region is damaged in a dose-dependent manner by repeated O3 exposures 0.2 ppm (Castleman et al., 1980; Van Bree et al., 2001). The damaged ciliated epithelium in the terminal and respiratory bronchioles is replaced by nonciliated Clara cells. In addition, interalveolar septa may become thickened by replacement of type 1 with type 2 cells and increased collagen synthesis leading to interstitial fibrosis. Chronic exposure of rats, using a diurnal O3 pattern to mimic typical urban ambient exposures (0.06  ppm for 13 h/day, 7 day/week, with a slow 9-h spike to 0.25 ppm 5 day/week) induced epithelial hyperplasia, fibroblast proliferation, deposition of collagen fibers, and increased density of basement membrane in the terminal bronchioles and proximal alveolar region (Chang et al., 1992). These changes were observed as early as the 13-week sacrifice. Bronchiolization of alveolar spaces was observed in a number of studies of rodents and primates chronically exposed to 0.5 to 1.0 ppm O3 (e.g. Fujinaka et al., 1985; Gross and White, 1987). Similarly, other investigators showed that with cyclic episodic exposures (1 ppm, 8 h/day for 5 days followed by 9 days of filtered air for 4 cycles),

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terminal bronchiolar remodeling, including thickened centriacinar interstitium and increased cellularity of airway epithelium, was cumulative while acute epithelial inflammation became attenuated with each consecutive episodic exposure (Schelegle et al., 2003). Increased antioxidant enzyme levels (e.g. superoxide dismutase) occur in regions most susceptible to O3 injury at prolonged exposures as low as 0.12  ppm. GSH reductase, GSH-S-transferase, and GSH peroxidase activity increased in a dose-dependent manner in centriacinar tissue following exposure to 0.12 ppm O3 in rats, mice, and guinea pigs for 3–90 days (Dormans et al., 1999; Plopper et al., 1994). In monkeys, the site-specific reduction in GSH pool following acute O3 exposures and the ability of different areas of the airway epithelium to replenish that pool, may be a factor in site-specific injury (Plopper et al., 1998).

19.6.5.3  Lung Host Defense AM are the primary cellular defense in the alveolar region, phagocytizing foreign particles and secreting chemicals that recruit and activate inflammatory cells in the lung. O3 exposure reduces phagocytic ability of AM. Introduction of infectious agents such as S. zoopidemicus, L. monocytogenes and S. aureus in mouse models following acute O3 exposure (0.1–0.8 ppm) results in decreased bacterial killing and increased morbidity and mortality due to infection (Cohen et al., 2001; Gilmour and Selgrade, 1993; Van Loveren et al., 1988). AM from humans exposed to 0.08 ppm O3 for 6.6 h while exercising showed a decrease in phagocytic activity of 25% relative to control, while mice exposed to 0.8 ppm for 3 h showed a 42% decrease in phagocytosis (Selgrade et al., 1995). Results from in vitro studies of human and murine AMs indicated similar sensitivity to decreased phagocytosis from O3 exposure. Considerable interspecies and interstrain variability in the impact of O3 on ROS production by AM is evident from a number of studies.

19.6.5.4  Effects of Ozone on Lung Development and Exacerbation of Asthma In a series of elegant studies by researchers at the Univ­ ersity of California, Davis, the effects of O3 exposure, alone or in combination with allergen, on infant rhesus monkey lung development included reduction in airway number, hyperplasia of the bronchiolar epithelium, increased mucous cell number, immune system dysfunction, and changes in innervations of the pulmonary epithelium (Plopper et al., 2007). Chronic cyclic O3 exposure (0.5 ppm, 8 h/day, 5 days on followed by 9 days of filtered air for 11 cycles) of infant rhesus monkeys induced biochemical and functional alterations [depleted proteoglycan and fibroblast growth factor-2 (FGF-2), altered FGF receptor-1], and thinning of the basement membrane zone (Evans et al., 2003). This tissue binds and releases growth

Chapter | 19  Volatile Organic Compounds from Pesticide Application and Contribution to Tropospheric Ozone

factors, is involved in cell–cell communication, and functions as a barrier. The alteration of FGF-2 signaling, important for regulating processes in the developing lung, may be associated with the reported O3-induced abnormal development of alveolar and bronchiolar regions in animal models. Cyclic repeated O3 exposure in the presence of house dust mite antigen (HDMA) also altered the development of the airways and enhanced the response to HDMA challenge in previously sensitized infant Rhesus monkeys (Schelegle et al., 2003). Cyclic O3 exposure increased airway hyperresponsiveness and serum histamine following allergen challenge, serum IgE, eosinophils in BALF, and the volume of mucous cells within the terminal bronchioles (Joad et al., 2006; Schelegle et al., 2003). Ozone exposure in the presence of allergen acts synergistically in the developing rhesus monkey lung to produce an allergicreactive phenotype airway. This has been postulated to occur in humans as well, and may underlie the correlation observed in epidemiologic studies between ozone exposure and new-onset asthma. Proper lung function in the mature animal is based in part on appropriately controlled innervation of the airways. Larson et al., (2004) demonstrated that cyclic O3 plus allergen exposure to infant monkeys reduces the density and pattern of innervation in mid-level pulmonary airways observed immediately after a 5-month regimen. In a follow-up study Kajekar et al. (2007) observed hyperinnervation and irregular distribution of neurons in mid-level airways 6 months following the cessation of ozone exposure. In this primate model, O3 and O3 plus allergen exposure lead to adaptive mechanisms that overcompensate for the initial disruption of airway innervation. Hyperinnervation is likely related to increased airway hyperresponsiveness. In a sensitized mouse model of asthma, short-term exposure to O3 (3 ppm for 2 h) exacerbates eosinophilic airway inflammation and hyperresponsiveness (Kierstein et al., 2008), both hallmarks of asthma. Relative to allergenchallenged mice breathing room air, O3 exposure greatly increased airway hyperresponsiveness (increased airway resistance in vivo, p 0.0001, and tracheal ring contractility in vitro, p  0.0006) and eosinophillia (measured as cells in BALF, p  0.0002). O3 exposure augmented release of pro-survival cytokines (IL-5, GM-CSF) as measured in BALF, inhibiting apoptosis and increasing survival of the eosinophils in the airway. Thus, one mechanism of asthma exacerbation appears to be sustained neutrophilic and eosinophilic airway inflammation by ozone, exacerbating the response to antigen.

19.6.6  Human Studies Investigators have used controlled exposures and observational epidemiology to study the effects of ozone exposure

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on respiratory health. The main focus of controlled human exposure studies has been the assessment of lung function before, during and after ozone exposure (Cal/EPA, 2005). Exposures are conducted in a chamber, often with intermittent exercise by the subjects to simulate outdoor activity. Controlled human exposures offer the advantage of wellcharacterized exposure and response, and avoid the need for cross-species extrapolation from animals to humans. However, limitations include small numbers of subjects, limited ability to examine respiratory tract tissues, and inability to study chronic exposure or subjects with moderate to severe lung disease. By contrast, observational epidemiology studies large numbers of individuals, both healthy and ill, in the real world, and uses powerful statistical methods to evaluate contributions of different pollutants to health outcomes. Many observational epidemiology studies have linked O3 exposure to a number of adverse health outcomes including exacerbation of asthma, induction of asthma, increased emergency room visits and hospitalizations for respiratory disease, and premature mortality.

19.6.6.1  Inflammation and Lung Function Several controlled exposure studies reported airway inflammation in O3-exposed subjects as increases in proinflammatory cytokines, protein, total cell count and specific cell numbers (AMs, PMNs, eosinophils) in BALF of ozone-exposed humans (Devlin et al., 1991; Seltzer et al., 1986). Lung function, measured using spirometry, captures airway status by volume and flow rates upon deep inhalation and forced expiration. Common metrics of lung function include: total amount of air in a forced exhalation (forced vital capacity, FVC); forced expiratory volume in 1 second, FEV1; the volume of air exhaled over the middle half (forced expiratory flow, FEF25–75) or at the latter part of expiration (FEF75); and peak expiratory flow rate (PEFR). FEV1 reflects large airways’ caliber, whereas FEF25–75 and FEF75 indicate small airways’ function. Single ozone exposures to between 0.5 and 1.0 ppm O3 in resting humans (e.g. Folinsbee et al., 1978; Silverman et al., 1976) reduced FVC, FEV1, FEF25–75. Increasing ventilation rate via exercise increased the magnitude of the response and decreased the concentration of O3 necessary to decrease lung function. Statistically significant decrements in lung function were observed in young healthy adults exercising at 0.12 ppm O3, the ozone standard at the time of the studies (e.g. Adams, 1986; Folinsbee et al., 1984; Gong et al., 1986; Linn et al., 1986; McKittrick and Adams 1995; Silverman et al., 1976). O3 exposure induced bronchoconstriction evidenced by decreased FEF25–75 in exercising healthy adults. Several investigators demonstrated a dose–response relationship for O3 and FEV1 in O3-exposed humans (6.6 h with intermittent exercise). Decrements in FEV1 were observed at 0.08 ppm in a dose-related fashion (Adams, 2002; Folinsbee et al., 1988,

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Horstman et al., 1990). Exposure to 0.04 ppm in Adams (2002) did not result in a significant decrease in group mean FEV1, although some individuals experienced significant decreases in FEV1. The current 8-h average NAAQS for O3 of 0.075 ppm is primarily based on the results of the controlled exposure studies demonstrating effects at 0.08 ppm for 6.6 h to individuals exercising intermittently. There is substantial intersubject variability in O3induced lung function among healthy adults (Folinsbee et al., 1978; Holz et al., 1999; McDonnell et al., 1993). Responsiveness is reproducible for each subject implying the role of innate characteristics. Specific gene loci in humans associated with O3-induced changes in lung function include TNF-, and enzymes involved in generation of ROS (NQO1) and detoxification of ROS (glutathione-Stransferase M1 and P1; heme oxygenase-1) (Alexeeff et al., 2008; Yang et al., 2008). Epidemiologic field studies have clearly demonstrated that ambient O3 is associated with reversible decrements in lung function, particularly in asthmatics (Linn et al., 1996; Mortimer et al., 2002; Neas et al., 1999; Ross et al., 2002). Significant associations are reported for PEFR and FEV1 in the afternoon and same-day or several day average O3 concentrations Attained lung function at maturity (age about 18 years) is decreased by O3 exposure in childhood. Kunzli et al. (1997) report a statistically significant negative association between lifetime O3 exposure and FEF25–75 in college freshman who had either lived in Los Angeles (median exposure of 51.5 ppb) or the San Francisco Bay Area (median exposure of 22.5 ppb) as children. Another investigation extended this study (Tager et al., 2005) and found similar results with a larger group (n  255). In both studies, the effects of O3 were independent of exposure to PM2.5 and NO2. A study conducted on 520 Yale freshman (Galizia and Kinney, 1999) also demonstrated significant reductions in FEF25–75 and in FEF75 in male students who had spent 10 years residing in locations with high O3 (peak 1-h O3 concentrations 80 ppb) compared with students residing in areas with lower exposures. In addition, statistically significant elevated odds ratios were noted for wheezing, and respiratory symptoms for long-term O3 exposures in childhood. These studies provide evidence that O3 impacts human lung development at the level of the small airways. Since the lungs have stopped growing by age 18 years, this effect is irreversible.

19.6.6.2  Asthma Exacerbation and Induction Controlled exposure studies indicate asthmatics are a sensitive population to O3 exposure. Asthmatics exposed to O3 experience wheezing, larger decreases in FEV1 and greater inflammatory reaction than nonasthmatics, exacerbated eosinophillic inflammatory response (Horstman et al., 1995; Kehrl et al., 1999; Scannell et al., 1996; Vagaggini

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et al., 2002), and increased sensitivity to aeroallergens (Chen et al., 2004). Many, though not all, observational epidemiologic studies find an association between ambient O3 exposure and exacerbation of asthma, in both children and adults (symptoms and medication use), that is independent of other air pollutants (e.g. Gent et al., 2003; Ross et al., 2002). Mortimer et al. (2000) report an association between O3 exposure and increased symptoms in 4- to 9-year-old children (n  846) with asthma (n  846). Gent et al. (2003) note significant associations in medicated asthmatics between symptoms and ambient O3 at 8-h average O3 levels 63.3 ppb, levels commonly encountered and below the NAAQS. A study of 138 children in Los Angeles reports elevated odds ratio for extra asthma medication use (Ostro et al., 2001). Other studies report significant associations between cough, phlegm, and difficulty breathing and O3 in asthmatic children between 6 and 13 years of age (Gold et al., 1999; Romieu et al., 1997). O3 exposure may also induce new-onset asthma in children. In a longitudinal cohort of children living in Southern California, children playing three or more sports in high O3 communities were found to be 3.3-fold more likely to develop new asthma than those playing no sports (McConnell et al., 2002). This was not observed in the lower O3 communities. Playing outdoor sports would result in higher O3 exposures.

19.6.6.3  Hospital Admissions and Emergency Room Visits for Respiratory Disease Many studies report positive associations between hospital admissions and emergency room visits for respiratory diseases and air pollution. Using hospital admissions data from 16 Canadian cities, Burnett et al. (1997) found a significant positive association between respiratory hospital admissions and previous day’s 1-h maximum O3 concentrations that was independent of co-pollutants. Control outcomes unrelated to the respiratory tract (e.g. digestive system) were not associated with ozone, indicating a specific effect of O3 on respiratory conditions. Asthma hospital admissions for children in Seattle were also correlated to O3 concentrations (Sheppard et al., 1999). Results from an analysis of six European cities indicated strong and consistent O3 effects on unscheduled hospital admissions for chronic obstructive pulmonary disease (COPD) (Anderson et al., 1997). Two separate analyses of a large dataset from Toronto, Canada, reported statistically significant daily increases in respiratory hospitalizations for all ages (Burnett et al., 1999) and for persons under age 2 years (Burnett et al., 2001). A study of data from 36 U.S. cities confirmed that increases in COPD and pneumonia hospitalizations were significantly associated with O3 exposure during the warm season (Medina-Roman et al., 2006). These studies demonstrated O3 effects that were not confounded by other air pollutants.

Chapter | 19  Volatile Organic Compounds from Pesticide Application and Contribution to Tropospheric Ozone

Friedman et al. (2001) evaluated emergency room (ER) visits, urgent care center visits, and hospital admissions for asthma in children 4 weeks before, during, and 4 weeks after the Atlanta Olympic Games in 1996, to evaluate the impact of a decrease in air pollutants due to actions taken to reduce traffic congestion. They found a reduction in asthma acute care events during the games, ranging from 11.1% to 44.1% (using four different databases), that correlated with a reduction in ozone and traffic. By comparison, changes in nonasthma acute care events ranged from 1.0% to 3.1%. Peak daily O3 concentrations decreased 28% from 81 ppb to 59 ppb during the Olympic games.

19.6.6.4  Mortality A number of studies have reported associations between O3 levels and mortality in adults. For example, Bell et al. (2006) analyzed O3 and mortality data encompassing the time period 1987–2000 from 98 U.S. cities and found that daily ozone levels are significantly associated with increases in the number of deaths on average across these cities. A recent report from the National Academy of Sciences, which evaluated the studies on O3 and premature mortality, concluded that short-term exposure to O3 is likely to contribute to premature deaths (NAS, 2008). The most susceptible individuals are those with pre-existing lung and heart disease.

Conclusion Volatile organic compounds (VOC) are organic chemicals that when released into the atmosphere can react with sunlight and nitrogen oxides (NOx) to form tropospheric (ground-level) ozone. Two general classes of pesticide products contribute the vast majority of pesticidal VOC emissions: fumigants and emulsifiable concentrates (EC). The relatively widespread use, high application rates, and high volatility underlie the large contribution of fumigants to the VOC problem from pesticide applications. The composition of EC are usually dominated by hydrocarbon solvents and carriers, which are sufficiently volatile to contribute to VOC emissions. The primary target organ for O3 health effects is the respiratory organs. O3 induces lipid peroxidation, a free radical chain reaction involving membrane lipids. Inflammation, which underlies the respiratory toxicity of ozone, involves a cascade of biochemical events following injury, including the release of pro-inflammatory cytokines by epithelial cells and macrophages, and influx of inflammatory cells into the alveolar and bronchiolar interstitium. Enhancement of tissue injury also occurs from release of proteolytic enzymes and reactive oxygen species (ROS) from inflammatory cells, resulting in additional oxidative damage to membrane lipids and proteins. Both animal and

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human studies demonstrate deficits in lung function following ozone exposure. Animal studies indicate structural changes in the lung following exposure during postnatal development. Ozone can act to exacerbate asthma and may be involved in induction of asthma in children. Recent studies implicate ozone in mortality from exacerbation of cardiovascular and pulmonary disease. Thus, control of tropospheric ozone is a major concern of regulatory agencies. Reduction of reactive VOC components of pesticide applications is important for the control of ozone in agricultural areas.

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Guth, D. J., Warren, D. L., and Last, J. A. (1986). Comparative sensitivity of measurements of lung damage made by bronchoalveolar lavage after short-term exposure of rats to ozone. Toxicology 40, 131–143. Haskell, D. (1995). “Potential Occupational and Non-occupational Exposure to Carbon Disulfide from Proposed Enzone Label Amendments to Allow Application with Various Types of Low-volume Irrigation Systems and Flood Irrigation,” Memorandum to G. Varnado dated August 15, 1995. Worker Health and Safety Branch. Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, CA 95814. Holz, O., Jorres, R. A., Timm, P., Mucke, M., Richter, K., Koschyk, S., and Magnussen, H. (1999). Ozone-induced airway inflammatory changes differ between individuals and are reproducible. Am. J. Respir. Crit. Care Med. 159, 776–784. Horstman, D. H., Folinsbee, L. J., Ives, P. J., Abdul-Salaam, S., and McDonnell, W. F. (1990). Ozone concentration and pulmonary response relationships for 6.6-hour exposures with five hours of moderate exercise to 0.08, 0.10, and 0.12 ppm. Am. Rev. Respir. Dis. 142, 1158–1163. Horstman, D. H., Ball, B. A., Brown, J., Gerrity, T., and Folinsbee, L. J. (1995). Comparison of pulmonary responses of asthmatic and nonasthmatic subjects performing light exercise while exposed to a low level of ozone. Toxicol. Ind. Health 11, 369–385. Joad, J. P., Kott, K. S., Bric, J. M., Peake, J. L., Plopper, C. G., Schelegle, E. S., Gershwin, L. J., and Pinkerton, K. E. (2006). Structural and functional localization of airway effects from episodic exposure of infant monkeys to allergen and/or ozone. Toxicol. Appl. Pharmacol. 214, 237–243. Johnson, B. (2006). “Calculation of EP Factors for 1,3-D for Five Areas for Periods from May 1 through October 31,” Memorandum to Randy Segawa dated November 30, 2996 http://www.cdpr.ca.gov/docs/emon/ pubs/ehapreps/analysis_memos/1903_3_apdx.pdf. Environmental Monitoring Branch, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, CA 95812–4015. Johnson, B., Barry, T., and Wofford, P. (1999). “Workbook for Gaussian Modeling Analysis of Air Concentration Measurements,” Department of Pesticide Regulation Notice dated February 20, 1998 EH99–03. http://www.cdpr.ca.gov/docs/emon/pubs/ehapreps/eh9903.pdf. Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, CA 95814. Kajekar, R., Pieczarka, E. M., Smiley-Jewell, S. M., Schelegle, E. S., Fanucchi, M. V., and Plopper, C. G. (2007). Early postnatal exposure to allergen and ozone leads to hypperinnervation of the pulmonary epithelium. Respir. Physiol. Neurobiol. 155, 55–63. Kehrl, H. R., Peden, D. B., Ball, B., Folinsbee, L. J., and Horstman, D. (1999). Increased specific airway reactivity of persons with mild allergic asthma after 7.6 hours of exposure to 0.16 ppm ozone. J. Allergy Clin. Immunol. 104, 1198–1204. Kierstein, S., Krystka, K., Sharma, S., Armani, Y., Salmon, M., Panettieri, R. A. Jr., Zangrilli, J., and Haczku, A. (2008). Ozone inhalation induces exacerbation of eosinophillic airway inflammation and hyperresponsiveness in allergen-sensitized mice. Allergy 63, 438–446. Kleinman, M. T., Mautz, W. J., and Bjarnason, S. (1999). Adaptive and non-adaptive responses in rats exposed to ozone, alone and in mixtures, with acidic aerosols. Inhal. Toxicol. 11, 249–264. Kunzli, N., Lurmann, F., Segal, M., Ngo, L., Balmes, J., and Tager, I. B. (1997). Association between lifetime ambient ozone exposure and pulmonary function in college freshmen—results of a pilot study. Environ. Res. 72, 8–23.

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Larson, S. D., Schelegle, E. S., Walby, W. F., Gershwin, L. J., Fannucchi, M. V., Evans, M. J., Joad, J. P., Tarkington, B. K., Hyde, D. M., and Plopper, C. G. (2004). Postnatal remodeling of the neural components of the epithelial-mesenchymal trophic unit in the proximal airways of infant rhesus monkeys exposed to ozone and allergen. Toxicol. Appl. Pharmacol. 194, 211–220. Last, J. A., Greenberg, D. B., and Castleman, W. L. (1979). Ozoneinduced alterations in collagen metabolism of rat lungs. Toxicol. Appl. Pharmacol. 51, 247–258. Last, J. A., Gelzleichter, T., Harkema, J., Parks, W. C., and Mellick, P. (1993). Effects of 20 months of ozone exposure on lung collagen in Fischer 344 rats. Toxicology 84, 83–102. Linn, W. S., Avol, E. L., Shamoo, D. A., Spier, C. E., Valencia, L. M., Venet, T. G., Fischer, D. A., and Hackney, J. D. (1986). A dose– response study of healthy, heavily exercising men exposed to ozone at concentrations near the ambient air quality standard. Toxicol. Ind. Health 2, 99–112. Linn, W. S., Shamoo, D. A., Anderson, K. R., Peng, R. C., Avol, E. L., Hackney, J. D., and Gong, H. (1996). Short-term air pollution exposures and responses in Los Angeles area schoolchildren. J. Expo. Anal. Environ. Epidemiol. 6, 449–472. McConnell, R., Berhane, K., Gilliland, F., London, S. J., Islam, T., Gauderman, J. W., Avol, E., Margolis, H. G., and Peters, J. M. (2002). Asthma in exercising children exposed to ozone: a cohort study. Lancet 359, 386–391. McDonnell, W. F., Muller, K. E., Bromberg, P. A., and Shy, C. M. (1993). Predictors of individual differences in acute response to ozone exposure. Am. Rev. Respir. Dis. 147, 818–825. McKittrick, T. and Adams, W. C. (1995). Pulmonary function response to equivalent doses of ozone consequent to intermittent and continuous exercise. Arch. Environ. Health 50, 153–158. Medina-Roman, M., Zanobetti, A., and Schwartz, J. (2006). The effect of ozone and PM10 on hospital admissions for pneumonia and chronic obstructive pulmonary disease: a national multicity study. Am. J. Epidemiol. 163, 579–588. Miller, P. D., Gordon, T., Warnick, M., and Amdur, M. O. (1986). Effect of ozone and histamine on airway permeability to horseradish peroxidase in guinea pigs. J. Toxicol. Environ. Health 18, 121–132. Mortimer, K. M., Neas, L. M., Dockery, D. W., Redline, S., and Tager, I. B. (2002). The effect of air pollution on inner-city children with asthma. Eur. Respir. J. 19, 699–705. NAS (National Academy of Sciences) (2008). “Estimating Mortality Risk Reduction and Economic Benefits from Controlling Ozone Air Pollution,” Committee on Estimating Mortality Risk Reduction Benefits from Decreasing Tropospheric Ozone. National Research Council. National Academy Press. Neas, L. M., Dockery, D. W., Koutrakis, P., and Speizer, F. E. (1999). Fine particles and peak flow in children: acidity versus mass.. Epidemiology 10, 550–553. NRC (1991). “Rethinking the Ozone Problem in Urban and Regional Air Pollution,” National Research Council Committee on Tropospheric Ozone Formation and Measurement. National Academy Press, Washington, DC. Oosting, R. S., van Golde, L. M., Verhoef, J., and van Bree, L. (1991). Species differences in impairment and recovery of alveolar macrophage functions following single and repeated ozone exposures. Toxicol. Appl. Pharmacol. 110, 170–178. Oshima, R. (1998). “Notice of Public Workshops to Identify Voluntary as Well as Enforceable Options for Reducing Volatile Organic

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Compound Emissions from Agricultural and Commercial Structural Use of Pesticides,” Department of Pesticide Regulation Notice dated February 20, 1998. Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, CA 95814. Ostro, B., Lipsett, M., Mann, J., Braxton-Owens, H., and White, M. (2001). Air pollution and exacerbation of asthma in AfricanAmerican children in Los Angeles. Epidemiology 12, 200–208. Pino, J. A. and Barry, T. A. (1992). “Estimation of Volatile Emission Potential of Pesticides by Thermogravimetry,” Study Protocol. Environmental Monitoring Branch http://www.cdpr.ca.gov/docs/ emon/pubs/protocol/tga_protocol.pdf. Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, CA 95812–4015. Pino, J. A., Barry, T. A., ����������������������������������������������� and�������������������������������������������� Rose, J. E. (1996). Estimation of Volatile Emission Potential of Pesticides by Thermogravimetric Analysis. Organic Methods Poster Session. Poster 09-C-001. In: “The 110th AOAC International Annual Meeting and Exposition. September 8– 12, 1996. Orlando, Florida.” Final Program p. 73. Plopper, C. G., Duan, X., Buckpitt, A. R., and Pinkerton, K. E. (1994). Dose-dependent tolerance to ozone. IV. Site-specific elevation in antioxidant enzymes in the lungs of rats exposed for 90 days or 20 months. Toxicol. Appl. Pharmacol. 127, 124–131. Plopper, C. G., Hatch, G. E., Wong, V., Duan, X., Weir, A. J., and Tarkington, B. K. (1998). Relationship of inhaled ozone concentration to acute tracheobronchial epithelial injury, site-specific ozone dose, and glutathione depletion in rhesus monkeys. Am. J. Respir. Cell Mol. Biol. 19, 387–399. Plopper, C. G., Smiley-Jewell, S. M., Miller, L. A., Fanucchi, M. V., Evans, M. J., Buckpitt, A. R., Avdalovic, M., Gershwin, L. J., Joad, J. P., Kajekar, R., Larson, S., Pinkerton, K. E., Van Winkle, L. S., Schelegle, E. S., Pieczarka, E. M., Wu, R., and Hyde, D. M. (2007). Asthma/allergic airways disease: does postnatal exposure to environmental toxicants promote airway pathobiology?. Toxicol. Pathol. 35, 97–110. Romieu, I., Meneses, F., Ruiz, S., Huerta, J., Sienra, J. J., White, M., Etzel, R. A., and Hernandez, M. (1997). Effects of intermittent ozone exposure on peak expiratory flow and respiratory symptoms among asthmatic children in Mexico City. Arch. Environ. Health 52, 368–376. Ross, M. A., Persky, V. W., Scheff, P. A., Chung, J., Curtis, L., Ramakrishnan, V., Wadden, R. A., and Hryhorczuk, D. O. (2002). Effect of ozone and aeroallergens on the respiratory health of asthmatics. Arch. Environ. Health 57, 568–578. Scannell, C., Chen, L., Aris, R. M., Tager, I., Christian, D., Ferrando, R., Welch, B., Kelly, T., and Balmes, J. R. (1996). Greater ozone-induced inflammatory responses in subjects with asthma. Am. J. Respir. Crit. Care Med. 154, 24–29. Schelegle, E. S., Miller, L. A., Gershwin Laurel, J., Fanucchi, M. V., Van Winkle, L. S., Gerriets, J. E., Walby, W. F., Mitchell, V., Tarkington, B. K., Wong, V. J., Baker, G. L., Pantle, L. M., Joad, J. P., Pinkerton, K. E., Wu, R., Evans, M. J., Hyde, D. M., and Plopper, C. G. (2003). Repeated episodes of ozone inhalation amplifies the effects of allergen sensitization and inhalation on airway immune and structural development in Rhesus monkeys. Toxicol. Appl. Pharmacol. 191, 74–85. Selgrade, M. K., Cooper, K. D., Devlin, R. B., van Loveren, H., Biagini, R. E., and Luster, M. I. (1995). Immunotoxicity—bridging the gap between animal research and human health effects. Fundam. Appl. Toxicol. 24, 13–21.

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Seltzer, J., Bigby, B. G., Stulbarg, M., Holtzman, M. J., Nadel, J. A., Ueki, I. F., Leikauf, G. D., Goetzl, E. J., and Boushey, H. A. (1986). O3-induced change in bronchial reactivity to methacholine and airway inflammation in humans. J. Appl. Physiol. 60, 1321–1326. Sheppard, L., Levy, D., Norris, G., Larson, T. V., and Koenig, J. Q. (1999). Effects of ambient air pollution on nonelderly asthma hospital admissions in Seattle, Washington, 1987–1994. Epidemiology 10, 23–30. Silverman, F., Folinsbee, L. J., Barnard, J., and Shephard, R. J. (1976). Pulmonary function changes in ozone-interaction of concentration and ventilation. J. Appl. Physiol. 41, 859–864. Spurlock, F. (2002a). “Analysis of the Historical and Revised Base Year 1990 Volatile Organic Compound Emission Inventories,” Memorandum to Randy Segawa dated December 16, 2002. Environmental Monitoring Branch http://www.cdpr.ca.gov/docs/ emon/vocs/vocproj/base_year_inv.pdf. Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, CA 95812–4015. Spurlock, F. (2002b). “Methodology for Determining VOC Emission Potentials of Pesticide Products,” Memorandum to John S. Sanders dated January 7, 2002. Environmental Monitoring Branch. http:// www.cdpr.ca.gov/docs/emon/vocs/vocproj/intro.pdf. Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, CA 95812–4015. Spurlock, F. (2005). “Revisions to Procedures for Estimating Volatile Organic Compound Emissions from Pesticides,” Memorandum to John S. Sanders dated February 7, 2005. Environmental Monitoring Branch http://www.cdpr.ca.gov/docs/emon/vocs/vocproj/voc_calc_ revision020405.pdf. Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, CA 95812–4015. Spurlock, F. (2006). “2006 Revisions to Procedures for Estimating Volatile Organic Compound Emissions from Pesticides,” Memorandum to John S. Sanders dated July 18, 2006. Environmental Monitoring Branch http://www.cdpr.ca.gov/docs/emon/vocs/vocproj/voc_calc_ revision071805.pdf. Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, CA 95812–4015. Sterner-Kock, A., Kock, M., Braun, R., and Hyde, D. M. (2000). Ozoneinduced epithelial injury in the ferret is similar to nonhuman primates. Am. J. Respir. Crit. Care Med. 162(3 Pt 1), 1152–1156. Subramanian, P. L., Teesch, , and Thorne, P. S. (1996). Degradation of 3,5-dimethyl-tetrahydro-2H-1,3,5-thiadiazine-2-thione in aqueous aerobic media. Environ. Toxicol. Chem. 15, 503–513. Tager, I. B., Balmes, J., Lurmann, F., Ngo, L., Alcorn, S., and Kunzli, N. (2005). Chronic exposure to ozone and lung function in young adults.. Epidemiology 16, 751–759. Tepper, J. S., Costa, D. L., Lehmann, J. R., Weber, M. F., and Hatch, G. E. (1989). Unattenuated structural and biochemical alterations in the rat lung during functional adaptation to ozone. Am. Rev. Respir. Dis. 140, 493–501. U.S.EPA (2006). “Air Quality Criteria for Ozone and Related Photochemical Oxidants (Final),” EPA/600/R-05/004aF-cF, 2006. U.S. Environmental Protection Agency, Washington, DC. Vagaggini, B., Taccola, M., Cianchetti, S., Carnevalli, S., Bartoli, M. L., Bacci, E., Dente, F. L., Di Franco, A., Giannin, D., and Paggiaro, P. L. (2002). Ozone exposure increases eosinophillic airway response induced by previous allergen challenge. Am. J. Resp. Crit. Care Med. 166, 1073–1077. van Bree, L., Dormans, J. A., Boere, A. J., and Rombout, P. J. (2001). Time study on development and repair of lung injury following ozone exposure in rats. Inhal. Toxicol. 13, 703–718.

Chapter | 19  Volatile Organic Compounds from Pesticide Application and Contribution to Tropospheric Ozone

Van Loveren, H., Rombout, P. J., Wagenaar, S. S., Walvoort, H. C., and Vos, J. G. (1988). Effects of ozone on the defense to a respiratory Listeria monocytogenes infection in the rat. Suppression of macrophage function and cellular immunity and aggravation of histopathology in lung and liver during infection. Toxicol. Appl. Pharmacol. 94, 374–393. Wernimont, G. T. (1985). Use of Statistics to Develop and Evaluate Analytical Methods. In (W. Spendly, ed.), Arlington, VA: Association of Official Analytical Chemists.

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Yang, I. A., Fong, K. M., Zimmerman, P. V., Holgate, S. T., and Holloway, J. W. (2008). Genetic susceptibility to the respiratory effects of air pollution. Thorax 63, 555–563. Youden, W. J. and Steiner, E. H. (1975). “Statistical Manual of the Association of Official Analytical Chemists,” Association of Official Analytical Chemists, Arlington, VA.

Chapter 20

Regulatory Aspects of Acute Neurotoxicity Assessment Sandra L. Allen Regulatory Science Associates, Dunoon, Argyll, United Kingdom

20.1  Introduction Neurotoxicity occurring as a consequence of a single exposure to a substance is a serious health concern, and incidents of human poisoning have occurred since antiquity. Historically, the majority of cases are due to exposure to plant and animal toxins, such as snake and spider bites, scorpion stings, and consumption of poisonous fish (Goonetilleke and Harris, 1999) and unripe fruit (Sherratt, 1995), as well as environmental toxins such as carbon monoxide (Prockop and Chichkova, 2007), mercury (Ekino et al., 2007), and alcohol (Harper, 2007). Many synthetic compounds, including therapeutic drugs, abused drugs (Devlin and Henry, 2008), and industrial solvents (Dick, 2006), are also known to have acute neurotoxic effects. Most insecticides have been developed for their neuroactive properties, albeit in insects. Consequently, assessment of neurotoxic potential of pesticides is an important part of the hazard and risk assessment processes. During the development of any new pesticide, a wide range of studies are performed to establish margins of safety for both occupational and environmental exposures. Typically, these studies are performed in laboratory animals and will comply with international regulatory guidelines. Regulatory studies are designed to allow as complete an evaluation as possible of any pharmacological or toxicological effects that may impact on human health, and since the late 1990s they have been updated to ensure that neurological effects are adequately assessed. This chapter discusses methods used to assess neurotoxicity in laboratory animals, regulatory toxicity studies, and specifically tests that are used to assess acute neurotoxic potential of pesticides.

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20.2  Toxicological effects following acute exposures to pesticides Acute exposure to pesticides, in the context of the risk assessment process, refers to human encounters that occur in the course of 1 day or less. The consequences of the exposures can occur immediately (acute toxic effects) or be delayed, and the effects may be transient, permanent, or even lethal. Testing methodology needs to be able to detect and characterize all such effects sufficiently to allow for their safe use. Therefore, studies include a variety of techniques (behavioral evaluation, neurochemical measurements, electrophysiology parameters, and histopathology) and assess animals from the point of exposure to a time that allows for detection of recovery from transient effects as well as onset of delayed effects.

20.3  Methodology for assessing neurotoxicity 20.3.1  Behavioral Methods Behavioral changes can be sensitive indicators of disturbed function of the nervous system because they may be observed earlier and/or at doses lower than demonstrable clinical symptoms or structural lesions (Alder and Zbinden, 1977, 1983; Broxup et al., 1989; Schulze and Boysen, 1991; Walsh and Chrobak, 1987). However, due to the functional reserve capacity of the nervous system, there is the possibility that some structural loss may occur in the nervous system while the animal remains functionally normal (Mitchell and Tilson, 1982).

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Not all behavioral changes necessarily represent the specific action of a chemical on the nervous system but can be indirect consequences of effects on other physiological systems. Many behavioral tests are affected by changes in nonneural organs and by dietary restriction, hormonal state, fatigue, motivation, age, housing conditions, as well as basic experimental design considerations such as strain and diet [European Centre for Ecotoxicology and Toxicology of Chemicals (ECETOC), 1992; Slikker et al., 2005]. Behavioral methods can be used to measure a wide variety of sensory, motor, cognitive, and autonomic functions (Tilson and Harry, 1992). Because most behaviors require the integrated activity of many components of the nervous system, many methods can provide information for more than one category of function. Interpreting results of any single behavioral test may be problematic, and interpretation is most effective when combined with additional behavioral methods or methods from other disciplines (i.e., electrophysiology, neurochemistry, and neuropathology). Behavioral tests are generally quantitative and noninvasive; thus, the same animal can be tested repeatedly during a toxicity study to provide detailed information about the presence or absence of effects, their severity, the time of onset, and duration or recovery. Some tests require special equipment and expertise, which can influence the numbers of animals that can be assessed concurrently and affect study scheduling.

Behavioral test methods have been reviewed and guidance on the choice of test methods to assess specific endpoints published many times [Annau, 1987; Cory-Slechta, 1989; Maurissen, 1995a; Organization for Economic Co-operation and Development (OECD), 2004; Tilson, 1990, 1997; World Health Organization (WHO), 1986, 2001]. Examples of commonly used test methods, the function primarily assessed by the test, and agents known to affect the measure are summarized in Table 20.1. Data from studies on sensory function need to be interpreted carefully because the measured endpoint is typically a motor response to a sensory stimulus. Thus, it is important that control measures are included to ensure that the change in response to the sensory stimulus is not due to motor dysfunction or an effect on the motivation to respond to the stimulus. Learning and memory are theoretical concepts that are sometimes difficult to separate experimentally, although some tests are designed to emphasize one or the other. Alterations in learning and memory must be inferred from changes in behavior. Many of the commonly used tests of learning and memory require that animals be trained to make a response to receive positive or avoid negative reinforcement. To conclude that an alteration in behavior results from a change in learning and memory, all other causes (e.g., effects on motor function, sensory perception, or motivation) must be excluded. Some cognitive tests are

Table 20.1  Tests Commonly Used to Measure Behavior and Example Agents Active in Animals Function

Tests

Agents

Weakness

Grip strength, landing foot splay

Acrylamide, 2,4,-D, chlordiazepoxide

Incoordination

Rotating rod (Rotorod, Accelerod)

Acrylamide, ethanol, IDPN

Locomotion

Motor activity

Amphetamine, chlorpromazine, carbaryl, scopolamine

Tremor

Rating scales, spectral analysis

Chlordecone, DDT, organochlorines, pyrethroids

Audition

Acoustic startle response, prepulse inhibition, auditory discrimination procedures

Antibiotics, DDT, pyrethroids

Vision

Discrimination task

LSD, methylmercury

Somatosensory

Discrimination task

Acrylamide, triethyltin

Pain perception

Tail flick test, hot plate test

Carbaryl, methylmercury, opiates

Active or passive avoidance, maze learning, swimming mazes, conditioned discrimination (matching to sample, repeated acquisition), eye blink conditioning, taste aversion, schedule-controlled operant behavior

Amphetamine, carbaryl, chlordecone, DDT, lead, lindane, scopolamine, triakyltins

Neuromuscular

Sensory

Cognitive Cognitive functions

Chapter | 20  Regulatory Aspects of Acute Neurotoxicity Assessment

capable of measuring learning as a function of a few trials, whereas others require several training trials over days or weeks. Some procedures assess short-term memory, whereas reference memory can be assessed in others (e.g., the Morris water maze). It is important to note that different tests assess different forms of learning or memory, and it is likely that any one chemical may show effects in some tests and not others depending on the procedure that is used. The choice of behavioral test depends on the purpose of the study; some tests may be simple to perform but lack sensitivity, whereas others are much more sensitive but are complex and time-consuming. However, complex tests are not necessarily more sensitive at detecting neurotoxicity. For example, a comparison of the relative sensitivity of a functional observation battery (FOB), motor activity (MA), and schedule-controlled operant behavior (SCOB) indicated that FOB was as sensitive as, or more sensitive than, MA or SCOB in detecting treatment-related effects (Moser and MacPhail, 1990). Simple observation of behavior is routinely included in all standard regulatory toxicity studies. More structured observations such as a FOB are also relatively easily included, and most regulatory agencies [e.g., OECD and the U.S. Environmental Protection Agency (EPA)] have updated protocols for many repeat-dose studies in recent years in order to include these types of measures. However, there are several problems related to analysis and interpretation

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of neurobehavioral screening data (Slikker et al., 2005; Tilson and Moser, 1992). Most screening batteries consist of several tests that yield different types of data that are each analyzed by different statistical methods. Each measure in the battery can be viewed as a unique endpoint, and because there are multiple tests in the battery, some statistically significant changes might occur just by chance (Type I error). This situation is compounded by the very large amounts of data in most screening experiments. The overall goal is to provide a robust and biologically plausible interpretation of the data using statistical analysis as a tool to support the data evaluation. It is essential that statistical analyses and significance criteria are established prior to generation of the data (Maurissen, 1995b; Muller et al., 1984), although multiple approaches are likely to be appropriate. Statistical considerations for analysis of behavioral test methods have been reviewed (Holson et al., 2008; Slikker et al., 2005).

20.3.2  Electrophysiological Techniques Electrophysiological techniques measure the electrical potentials of impulse transmission in the nervous system and thus reflect the function of neurons. Electrical potentials can be recorded in specific areas of the central nervous system (CNS) or peripheral nervous system (PNS) in vivo or from in vitro preparations (see Table 20.2 for a summary of methods and examples of agents known to have effects).

Table 20.2  Electrophysiological Methods Used to Determine Neurotoxicity and Example Agents Active in Animals Test

Measurement

Agents

Visual evoked potentials (flash evoked potentials and pattern reversal evoked potentials) (VEPs)

Record of integrated cortical response to a flash of illumination or patterns of visual stimuli reflecting activity in the visual pathway

Triethyltin, carbon monoxide

Auditory evoked potentials [AEPs; and brainstem auditory evoked response (BAER)]

Record of integrated cortical response to auditory stimuli (tones or clicks) reflecting electrical activity in auditory pathway

Aminoglycosides, toluene, styrene

Somatosensory evoked potentials (SSEPs)

Record of integrated responses of somatosensory pathway in response to mechanical, electrical, thermal, or proprioceptive stimuli; recorded from cerebellum or cortical surface

Acrylamide, carbon disulfide, lead

Peripheral nerve (motor or sensory) evoked potential, nerve conduction velocity (NCV)

Record of action potential recorded using electrodes following electrical stimulation of nerve

Acrylamide, carbon disulfide, hexachlorophene, methylmercury

Electromyography (EMG)

Record of electrical potential in muscle fibers when they contract in response to mechanical or electrical stimuli

Manganese, organophosphates

Electroretinography (ERG)

Record of electrical responses of various cell types in the retina, including the photoreceptors, inner retinal cells, and the ganglion cells in response to flash of illumination or patterns of visual stimuli

Methanol

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Advantages of electrophysiological methods to the neurotoxicologist include the relative ease with which the data can be quantified, analyzed, and standardized as well as the large amount of electrophysiological data that can be collected quickly. Also, some techniques are noninvasive and allow monitoring of progression and/or recovery of functional disturbance. However, electrophysiological methods require specialist knowledge and equipment, are not always compatible with standard toxicity studies due to the need for electrode implantation [visual evoked potentials (VEPs), brainstem auditory evoked response (BAER), and somatosensory evoked potentials (SSEPs)], and are not commonly included in screening studies. Electrophysiological techniques are used extensively in human clinical neurology and are readily applied with minimal modification across species (Arezzo et al., 1985; Dyer, 1985; Mattsson et al., 1989; Rebert, 1983; Seppalainen, 1975). Most animal electrophysiology data are extrapolated easily to humans because these data are familiar to the medical community (Mattsson and Albee, 1988). In fact, the degree of comparability to humans typically is higher for electrophysiology tests than for most behavioral measures (Winneke, 1992). The electrophysiological method for a particular experiment must be appropriate to the question being asked. If one is interested in overt changes, then macroelectrode procedures, such as electroencephalography or evoked potentials, may be adequate [Office of Technology Assessment (OTA), 1990]. More specific questions, such as whether the chemical acts on presynaptic receptors, specific ion channels, or sensory rather than motor nerves, demand more sophisticated experimental procedures (Atchison, 1988; Kerkut and Heal, 1981). The latter techniques, perhaps in in vitro preparations (Rowan, 1985), may provide specific information on the mechanism of neurotoxicity of a particular chemical.

20.3.3  Neurochemical Endpoints Neurochemical methods are used increasingly to investigate mechanisms of action of neurotoxic chemicals (Bondy, 1986; Costa, 1988; Silbergeld, 1987; WHO, 2001). Functioning of the nervous system depends on multiple neurochemical processes, and any chemical-induced change could potentially result in neurotoxicity (see Table 20.3 for effects that can be detected that may indicate evidence of neurotoxicity). Neurochemistry can be conducted on parts of peripheral nerves, the entire brain of animals, distinct brain structures obtained by dissecting whole brains, slices of whole brain or particular brain structures, as well as neurons or glial cells cultured in vitro. To increase the sensitivity of neurochemical methods, cells can be fractionated and particular cell organelles separated. Because neuronal lesions generally are limited to specific areas of the brain and often to specific types of

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Table 20.3  Biochemical Effects Indicative of Potential Neurotoxicity Alterations in synthesis, release, uptake, and degradation of neurotransmitters Alterations in second messenger-associated signal transduction Alterations in membrane-bound enzymes regulating neuronal activity Inhibition and aging of neuropathy target enzyme Increase in glial fibrillary acidic protein in adults Decreases in mRNA or protein synthesis Increased production of oxygen radicals Changes in energy-related functions Synthesis of heat shock proteins

neurons, the sensitivity of neurochemical measurements decreases with increasing volume of nerve tissue in a single assay. In contrast, the chance of missing an effect increases with decreasing total volume of tissue. The relationship between nervous system function and observations made in neural tissue extract in a test tube can be somewhat tenuous because the concentration of many endogenous substances or the activity of enzymes may change rapidly after death. Further problems arise due to the tendency of the nervous system to compensate for neuronal loss, such as by increasing turnover rate of transmitters or by up- or downregulation of receptors, but such changes are not indicative of a neurotoxic effect unless they are likely to have neurobehavioral, neuropathological, or neurophysiologic consequences. Neurochemical methods are not routinely used in standard toxicity studies with the exception of toxicity tests for organophosphate and carbamate pesticides that include measure of acetylcholinesterase and/or neuropathy target enzyme. An approach that overcomes some of the problems related to neurochemical assays is the combination of neurochemical determinations and histopathology (histochemical staining techniques). Such techniques are specific, sensitive, and have the advantage of showing the topographic distribution of findings (Krinke and Hess, 1981). Immunoassays have also been developed for neurotypic proteins such as the major intermediate filament protein in astrocytes, glial fibrillary acidic protein (GFAP). Immunoassay measures are better at quantifying changes in protein levels and providing dose–response information. Immunohistochemistry and tissue immunoassays have shown increased levels of GFAP in developing and mature animals exposed to chemicals known to damage the CNS (Brock and O’Callaghan, 1987; O’Callaghan, 1988; O’Callaghan and Miller, 1988).

Chapter | 20  Regulatory Aspects of Acute Neurotoxicity Assessment

20.3.4  Neuropathological Methods The morphological complexity of the nervous system must be taken into account in the application of pathological techniques for the assessment of the neurotoxic potential of chemicals in animals. The methods and factors that affect the ability to detect neuropathology have been reviewed (ECETOC, 1992; OECD, 2004; WHO, 1986) and practical guidance has been published (Fix and Garman, 2000). Standard pathological methods involving direct immersion fixation of brain, spinal cord, and peripheral nerves in 10% neutral buffered formalin are considered adequate for routine screening studies, although artifactual changes are likely to occur (Garman, 1990). More detailed examination of the nervous system usually requires perfusion fixation and special stains, along with other specialized procedures to define particular effects and avoid misinterpretation of artifacts (O’Donoghue, 1989; Mattsson et al., 1990). Perfusion fixation with formalin, paraformaldehyde, and/or glutaraldehyde is commonly used for specific investigations of neuropathological changes in the CNS or PNS (Krinke, 1989; Mattsson et al., 1990; O’Donoghue, 1989). Perfused tissue may be embedded in paraffin for routine light microscopy or postfixed in osmium tetroxide and processed for plastic-embedded semithin (1 or 2 m) sections and stained with toluidine blue. Nervous tissue fixed by perfusion is devoid of most artifacts associated with immersion fixation (Garman, 1990); nevertheless, perfusion may be associated with other artifacts, such as those caused by inadequate control of pressure, pH, or osmolarity of the fixative (Schultz and Karlsson, 1965). If immunohistochemical methods are to be used, then frozen sections may be required, which will usually necessitate the inclusion of additional animals in the study. The macroscopic complexity of the brain causes difficulty in tissue sampling, and to achieve a reasonably comprehensive survey of the CNS multiple coronal sections need to be examined at a number of levels. Between five and seven levels are commonly used and need to be standardized using landmarks to ensure consistency of areas evaluated between animals. Ultrathin sections of small blocks of perfused, plastic-embedded nervous tissue can be examined with a transmission electron microscope. Although electron microscopy is a powerful tool in mechanistic studies of neurotoxicology, its application is laborious (WHO, 1986), and due to the small tissue sample size used it will miss lesions unless tissue selection is accurate. Therefore, ultrastructural studies are generally confined to those protocols where there is a specific need to characterize neuropathological changes and answer mechanistic questions. Histological stains are available to help detect and characterize effects. New or improved stains and antibodies now available include lectin histochemistry to assess glial and microglial alterations (Fix et al., 1996), Fluoro-Jade

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to detect necrosis (Krinke et al., 2001; Schumed and Hopkins, 2000), or silver stains specific for neural degeneration (Switzer, 2000). Nerve fiber teasing (separation of perfusion-fixed peripheral nerve fibers embedded in epoxy resin) may provide a valuable means to characterize peripheral neuropathy (Krinke et al., 2000). Interpretation of pathological changes may be confounded by spontaneous background lesions that are common due to aging and trauma. The more common background lesions in the rat have been reviewed (McMartin et al., 1997; Mohr et al., 1994). Neuropathological lesions can be classified according to the site where they occur (Krinke, 1989; Spencer et al., 1980; WHO, 1986) and should be described using standard terminology (McMartin et al., 1997). Morphometric (quantitative) evaluation may assist in the detection of a treatment-related effect and in the interpretation of treatment-related differences in brain weight or morphology (De Groot et al., 2005a,b). Linear or areal measurements allow for lengths or thickness of structures to be measured. Quantification of the total number of neurons or other cell types, volume of brain regions, surface areas, and size of neurons or neuronal nuclei can be performed using stereology (Scallet, 1995). Although neuropathology provides clearly interpretable data and high resolution (including single neurons and axons), the methods are limited to static evaluation of discrete sections and thus should be integrated with functional studies.

20.4  Standard acute toxicity studies Standard acute toxicity studies are designed to provide information to enable appropriate classification and labeling of the substance under test. Primarily they are focused on defining a median lethal dose following exposure by the oral, dermal, and inhalation routes, and studies are conducted on both the active ingredient and the pesticide formulations. They use relatively small numbers of animals, at a limited number of dose levels, and typically the only measures are body weight, clinical signs, and mortality. Rarely, morphological assessment of tissues is also performed on animals that survive to scheduled termination 2 weeks after the acute dose or that die or are killed as a consequence of acute toxicity during the course of the study. Nevertheless, standard studies can provide information on neurotoxic potential. Clinical observations included in standard toxicity study protocols usually are obtained by cage-side monitoring of animals as well as during handling at the time of dosing or body weight determination. Clinical observations may indicate changes in motor function (e.g., disturbances of gait and abnormal posture or muscle tone), level of arousal (e.g., hyperactivity and lethargy), autonomic functions (salivation, lacrimation, urination,

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and defecation), and psychological status (indicated by stereotyped behavior, aggression, biting, licking, and selfmutilation) or may indicate pharmacological effects (sedation and anesthesia). When there is a particular concern for neurotoxicity (e.g., based on structure–activity relationships, known mechanism of action, or results of preliminary studies), specific neurotoxicity studies should be considered. Regulatory studies specifically designed to investigate potential neurotoxicity have been issued only for industrial chemicals and pesticides (Table 20.4).

20.5  Regulatory neurotoxicity studies Guidelines were first introduced for organophosphorus chemicals in the early 1980s, with other study types being introduced in the mid-1980s, and they have been regularly updated. In addition, the U.S. EPA also issued guidance for acute (and repeat dose) comparative sensitivity studies for measuring cholinesterase (U.S. EPA, 2001). The data from these studies were intended to be used to define the comparative sensitivity of adults and young organisms to support risk assessments under the Federal Insecticide, Fungicide, and Rodenticide Act and the Food Quality Protection Act of 1996 (FQPA). All of these guidelines are publicly available and are described in the following sections.

Table 20.4  Acute Neurotoxicity Testing Guidelines Organization

Reference Title

JMAFF

2-1-7

Acute neurotoxicity study

JMAFF

2-1-8

Acute delayed neurotoxicity study

OECD

418

Delayed neurotoxicity of organophosphorus substances following acute exposure

OECD

424

Neurotoxicity study in rodents

OPPTS

870-6100

Acute and 28-day delayed neurotoxicity of organophosphorus substances

OPPTS

870-6200

Neurotoxicity screening battery

OPPTS

870-6500

Schedule-controlled operant behavior

OPPTS

870-6850

Peripheral nerve function

OPPTS

870-6855

Neurophysiology: Sensory evoked potentials

JMAFF, Japanese Ministry of Agriculture, Farming and Fisheries; OECD, Organization for Economic Co-operation and Development; OPPTS, Office of Prevention, Pesticides and Toxic Substances, U.S. Environmental Protection Agency.

20.5.1  Delayed Neurotoxicity of Organophosphorus Compounds Organophosphorus compounds (OPs) have diverse effects on the PNS and CNS due to their ability to inhibit acetylcholinesterase and/or neurotoxic esterase. Inhibition of acetylcholinesterase produces the signs seen in mammals following acute poisoning. The signs are related to excess acetylcholinesterase and the consequent overstimulation of all parts of the PNS and CNS that use acetylcholine as the neurotransmitter. Signs of acute poisoning usually occur within minutes to a few hours of exposure and include excessive urination, lacrimation, diarrhea, muscular twitching, weakness, and convulsions. With severe poisoning, death can occur, usually caused by respiratory paralysis (Lotti, 2001; O’Brien, 1960, 1967). Some OPs also cause inhibition of the membrane-bound protein neurotoxic esterase (NTE), which can result, after a single dose, in a delayed polyneuropathy. The physiological functions of NTE are unknown, and it is not clear how phosphorylation and aging of NTE leads to axonal degeneration (Lotti and Moretto, 2005). Clinical onset is delayed for approximately 1–3 weeks and mainly affects the lower limbs. Recovery can occur, but there is no specific treatment (Barrett et al., 1985) and permanent disability is common. Although OP neuropathy has been demonstrated in a number of species, there is considerable variation in susceptibility (Johnson, 1975). The female domestic hen (Gallus gallus domesticus) is classically the species of choice because the response is consistent and reproducible (Cavanagh, 1964a). Following a single dose of a neurotoxic OP, there is rapid inhibition of NTE that can be detected 1 or 2 days later in the in vitro assay of nervous tissue from dosed hens. The percentage inhibition, and thereby the degree of phosphorylation of NTE, is highly correlated with the initiation of OP delayed neuropathy (not all OPs that inhibit NTE cause neuropathy, but all those that cause neuropathy inhibit NTE). Approximately 1 week following a single dose, clinical signs of neuropathy first become apparent— the bird walks with an unsteady flat-footed gait. Depending on the dose, the signs become more severe until the bird is unable to stand and the weakness affects the wings, which may droop. The morphological pattern of OP distal axono­ pathy consists of symmetrical, distal axonal degeneration of ascending and descending nerve fiber tracts located in the CNS and PNS. Primarily, long, large-diameter fibers are affected (Bischoff, 1967, 1970; Cavanagh, 1964b; Prineas, 1969; Spencer and Schaumburg, 1976, 1978). In the PNS, the longer nerve fibers to the hindlimbs are affected before the shorter fibers to the forelimbs. Concurrently, the long spinal cord tracts, such as the dorsal columns (corticospinal and spinocerebellar tracts), show distal axonal degeneration. The degenerative change appears to move in a retrograde manner along the affected pathway with neuronal damage increasing in severity from proximal to distal regions.

Chapter | 20  Regulatory Aspects of Acute Neurotoxicity Assessment

The regulatory studies (Table 20.5) are designed to detect the functional, biochemical, and pathological deficits of OP-induced polyneuropathy. Thus, adult, laying hens receive a single (typically oral) dose of the test substance. Because the dose required to induce neuropathy is normally higher than that which causes cholinergic stimulation, the dose used should be equivalent to the median lethal dose or an approximate lethal dose, although doses higher than 2 g/kg are not required to be tested. If it is known from preliminary data that significant cholinergic signs are seen soon after dosing, then prophylaxis using atropine (20 mg/kg subcutaneously every 2 h)

Table 20.5  Study Designs for Acute Delayed Neurotoxicity of Organophosphorus Substances Species and age

Hens, 8–12 (14) monthsa

Group size

6 for biochemistry, 6 for pathology

Control groups

Vehicle control (6 for biochemistry, 6 for pathology)

Positive control

3 for biochemistry, 3 for pathology; recent historical data acceptable

Dose levels

One dose, single exposure

Route

Normally oral (by gavage); dermal acceptable if appropriate

Observation period

21 days after dosing

Observations and frequency

Signs of toxicity—daily Ataxia, paralysis on a 4-point scale—daily Forced activity (e.g., ladder climbing) outside of home cage—twice weekly Body weight—weekly

b

NTE, AChE

Control and treated groups: 3 hens at 24 h, 3 hens at 48 h (unless otherwise indicated) for NTE (brain and spinal cord) Positive control: 3 hens at 24 h

Neuropathology

Gross necropsy all animals including those moribund; at termination in situ fixation, myelin, and axon-specific stains

Tissues/sections

Cerebellum, medulla oblongata, spinal cord (three levels), peripheral nerves (proximal tibial nerve and distal branches)

a

 OECD, 8–12 months; OPPTS and JMAFF, 8–14 months.  AChE optional for JMAFF and OECD.

b

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may be appropriate to prevent acute cholinergic deaths. Nevertheless, acute toxic deaths are not uncommon and so the number of animals dosed has to be increased to ensure that a sufficient number survive for neuropathology (minimum of 6 required) and biochemical investigations (minimum of 6); groups of approximately 20 are often used. A control group of 12 animals (6 for neuropathology and 6 for biochemistry) and a positive control group (containing sufficient animals to allow 3 to be assessed for neuropathology on day 21 and at least 3 assessed for biochemistry) are also required. Positive control data do not need to be concurrent if “recent” background data are available. It is recommended that positive control data be updated whenever some essential element of the test has changed or the data are more than 3 years old. The most widely recognized and commonly used positive control agent is triortho-cresyl phosphate (TOCP). Hens are monitored daily for a period of at least 21 days for the onset, severity, and duration of clinical signs of toxicity. Specific attention is paid to signs of gait abnormality and/or paralysis. At least twice per week, animals are subjected to a period of forced locomotor activity during which a semiquantitative assessment of locomotor deficit should be used to grade ataxia. Use of a rating scale of at least 4 points as in Roberts et al. (1983) is recommended. At predetermined time points (usually 24 and 48 h following dosing), subsets of hens (3 per time point) are killed and the brain and lumbar spinal cord are removed and prepared for the in vitro determination of NTE (Johnson, 1977, 1982; Kayyali et al., 1991; Soliman et al., 1982; Sprague et al., 1981; Zech and Chemnitius, 1987). Sciatic nerve tissue may also be assayed for NTE (Carrera et al., 1994; Moretto et al., 1991; Tormo et al., 1993). The test method is a differential assay of the ability of neural tissue, following OP exposure, to hydrolyze a phenyl valerate substrate selectively. The principle of the assay is first to determine the amount of hydrolysis that occurs in the presence of a nonneurotoxic inhibitor, paraoxon (a), which is intended to occupy irrelevant sites, and second to determine the activity in the presence of paraoxon and a known neuropathic inhibitor, mipafox (b). NTE activity is the difference between a and b—that is, the proportion of activity inhibited only by mipafox. Thus, the “mipafox site” is already occupied following exposure to a neuropathic OP ester and the activity of b is therefore reduced. Measurement of acetylcholinesterase (AChE) activity (in the same tissue in which NTE is measured) is required by the Office of Prevention, Pesticides and Toxic Substances of the U.S. EPA [OPPTS; it is optional for OECD and the Japanese Ministry of Agriculture, Farming and Fisheries (JMAFF)]. The method used is based on the Ellman method (Ellman et al., 1961; Johnson and Russell, 1975) and uses acetylthiocholine hydrolysis and colorimetric techniques to determine the extent of AChE inhibition in blood and brain.

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At the end of the study, all survivors are given a macroscopic examination and samples of the nervous system are prepared for microscopic examination. Tissues are fixed by whole-body perfusion and representative samples of both CNS and PNS prepared (see Table 20.5). Sections are stained with appropriate myelin and axon-specific stains and examined at the light microscope level. Interpretation of the study should be in terms of the incidence, severity, and correlation of behavioral, biochemical, and histopathologic effects and any other observed effects in the treated and control groups. The conclusion that a chemical may produce delayed neuropathy should be based on at least two of three factors: (1) evidence of a clinical syndrome, (2) pathological lesions, and (3) NTE inhibition. NTE needs to be inhibited by at least 55–70% after acute exposure in order to initiate neuropathy (Ehrich et al., 1995).

20.5.2  Neurotoxicity Screening Battery/ Neurotoxicity Study A summary of the acute mammalian regulatory neurotoxicity studies as described by OECD (424 Neurotoxicity Study in Rodents), U.S. EPA (870-6200 Neurotoxicity Screening Battery), and JMAFF (2-1-7 Acute Neurotoxicity Studies) is given in Table 20.6. The designs of the studies are essentially identical. The guidelines allow for single- or repeatdose studies (up to 2 years in duration, although generally acute, single-dose studies and 90-day repeat-dose studies are performed). Studies are usually performed in the rat, although other species may be used if appropriate. The study incorporates an observation battery (FOB), assessment of locomotor activity, and detailed neuropathology, although other methodology (e.g., neurochemistry and electrophysiology) may be included to more fully characterize any effects provided inclusion does not compromise the basic study design. Observations are to be conducted blind, at specific times, and in a structured manner in the following order: observations in the home cage, while the rat is handled and held during removal from the cage, and while the rat is moving in a standard open arena and through manipulative tests. Specific measures to be included are listed in Table 20.7. Because most of the assessments of FOB are subjective, it is important that observers are carefully trained and that observations are performed using explicitly defined scales [usually described in standard operating procedures (SOPs)] that require the absence as well as the presence of each observation to be recorded. Examples of such scoring systems have been published (Haggerty, 1989; McDaniel and Moser, 1993; Moser et al., 1988; O’Donoghue, 1989). Careful adherence to SOPs is essential for these assessments because changes in procedural details such as animal handling, timing and methods of equipment cleaning,

Table 20.6  Acute Neurotoxicity Study Designs: JMAFF 2-1-7, OECD 424, and OPPTS 870-6200 Species and age

Young adult rat; 5–6 weeks olda/ weaning to 9 weeksb/at least 42 days oldc. Other species may be used if more appropriate.

Group size

At least 10 males and 10 females per group.

Control groups

Concurrent vehicle and/or untreated control required.

Positive control

Historic control data acceptable. Data should demonstrate sensitivity of methods and ability of methods to detect neurotoxicity.

Dose levels

At least three dose levels.

Route of exposure

Most appropriate based on likely human exposure, bioavailability, and practicality; typically gavage.

Observations

Body weight. Food consumption. Functional observation battery (detailed clinical observations and quantitative measurement of grip strength and landing foot splay) performed blind. Motor activity – with automated devices.

Frequency of observations

Prior to exposure and at estimated time of peak effect within 8 h of dosing, and 7 and 14 days after dosing.

Neuropathology

At least 5 males and 5 females per group. In situ perfusion fixation required. Special stains recommended. Representative samples of tissues to allow thorough examination of the nervous system.

a

 JMAFF acute neurotoxicity studies.  OECD neurotoxicity study.  OPPTS neurotoxicity screening battery.

b c

and changes in the test environment (e.g., lighting or noise levels) can all significantly affect the behavior of the test animals. Variability in test procedures can result in increased variability in test results, which decreases sensitivity of the test to detect the effect of the test substance on behavior. Because the standard study includes 80 animals and because of the time required to assess each animal, it is usually necessary to stagger dosing and evaluations over several days to control for the time of day. Replicated designs are recommended such that animals from all groups are tested in the same time frame on each day of testing.

Chapter | 20  Regulatory Aspects of Acute Neurotoxicity Assessment

Table 20.7  Measures for the Functional Observation Battery Assessment of signs of autonomic function, including but not limited to Ranking of the degree of lacrimation and salivation, with a range of severity scores from none to severe The presence or absence of piloerection and exophthalmus Ranking or count of urination and defecation, including polyuria and diarrhea Pupillary function such as constriction of the pupil in response to light or a measure of pupil size Degree of palpebral closure (e.g., ptosis) Description, incidence, and severity of any convulsions, tremors, or abnormal motor movements, both in the home cage and in the open field Ranking of the subject’s reactivity to general stimuli such as removal from the cage or handling, with a range of severity scores from no reaction to hyper-reactivity Ranking of the subject’s general level of activity during observations of the unperturbed subject in the open field, with a range of severity scores from unresponsive to hyperactive Descriptions and incidence of posture and gait abnormalities observed in the home cage and open field Ranking of any gait abnormalities, with a range of severity scores from none to severe Forelimb and hindlimb grip strength measured using an objective procedure Quantitative measure of landing foot splay Sensorimotor responses to stimuli of different modalities (e.g., pain perception and audition) Body weight Description and incidence of any unusual or abnormal behaviors; excessive or repetitive actions (stereotypies); emaciation; dehydration; hypotonia or hypertonia; altered fur appearance; red or crusty deposits around the eyes, nose, or mouth; and any other observations that may facilitate interpretation of the data Additional measures that may also be included Counting of rearing activity on the open field Ranking of righting ability Body temperature Excessive or spontaneous vocalizations Alterations in rate and ease of respiration (e.g., rales or dyspnea) Sensorimotor responses to visual or proprioceptive stimuli

Training of observers is most effectively done using positive control agents, although training videos and reference manuals for conducting an FOB have been produced. Periodic updating of training and a mechanism for

595

demonstrating interobserver reliability are also required because it is not always possible for the same observer to evaluate all animals at all time points in a single study. Quantitative measures included in the FOB are forelimb and hindlimb grip strength (Meyer et al., 1979) and landing foot splay (Edwards and Parker, 1977). Sensorimotor responses may be assessed qualitatively or quantitatively. The tail flick test (based on the method of D’Amour and Smith, 1941) is a simple and reliable measure of pain perception used in some laboratories. Locomotor activity is assessed with an automatic device capable of detecting both increases and decreases in activity. Each animal should be tested individually, and recording sessions need to be sufficiently long to ensure that motor activity approaches asymptotic levels by the last 20% of the session in control animals. Environmental conditions must be controlled and variations across treatment groups minimized. Therefore, treatment groups need to be counterbalanced across motor activity devices as well as time of day, and study design will be influenced by the number of devices available. The type of device to be used is not defined, but the most common ones are photodetectors, infrared detectors, or video imaging and assess activity in a novel environment (typically a square or rectangular cage similar to a home cage but without bedding or nesting material, although more complex arenas such as the figure-8 are also used). The duration of the test session required to meet the preceding criteria will vary depending on the type of device, shape or novelty of the arena, as well as the strain, sex, and age of animal, although durations of 30–90 min are typical. Positive control data are required to demonstrate the sensitivity and reliability of the activitymeasuring device and testing procedures. These data should demonstrate the ability to detect chemically induced increases and decreases in activity. Pharmacological agents such as amphetamine (increase in activity) or chlorpromazine (decrease in activity) are commonly used. At the end of the study, five animals per sex per group are killed using in situ perfusion fixation and a comprehensive evaluation of the nervous system is performed. A list of the commonly sampled tissues and sections examined is given in Table 20.8. Generally, the brain, spinal cord, eye, and skeletal muscle are embedded in paraffin and stained with hematoxylin and eosin, whereas the peripheral nerves are embedded in plastic and stained with toluidine blue. Additional, specialist stains may be used to investigate any effects seen. The evaluation of the study should take into account all preceding and/or concurrent toxicity data and correlation of functional and histopathological findings. Evidence of any dose–response relationship and the presence or absence, incidence, and severity of any neurotoxic effects need to be carefully considered. The evaluation should include appropriate statistical analyses, taking into account the concerns and issues discussed previously.

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Table 20.8  Representative Areas of the Nervous System for Histopathology Examination

Table 20.9  OPPTS 870-6500 Schedule-Controlled Operant Behavior

Brain

Species and age

Six or seven levels/coronal sections to include Forebrain

Young adult rat or mouse, at least 14 or 6 weeks old (respectively) prior to exposure. Other species may be used if more appropriate.

Center of the cerebrum (including a section through the hippocampus) Midbrain

Group size

8–10 animals of one sex per group.

Cerebellum

Control groups

Concurrent vehicle and/or untreated control required.

Positive control

Historic control data acceptable. Data should demonstrate sensitivity of methods and ability of methods to detect neurotoxicity.

Dose levels

At least three dose levels.

Pons Medulla oblongata Dorsal root ganglia and the dorsal and ventral root fibers Cervical region Lumbar region Spinal cord Cervical (at cervical swelling) – longitudinal and transverse sections Lumbar (at lumbar swelling) – longitudinal and transverse sections Eye, with optic nerve and retina

Highest dose does not need to exceed 2 g/kg. Route of exposure

Most appropriate based on likely human exposure, bioavailability, and practicality. Typically gavage.

Observations

Measurements of operant behavior for animals trained to respond under a schedule of reinforcement (e.g., fixed ratio or fixed interval) before and after exposure. Body weight.

Training

Animals need to be trained until they achieve stable performance. This should be at same time of day and typically will be 5 days a week for several weeks.

Frequency of observations

At time of estimated peak effect. Experimental session will typically have a duration of 30–40 min.

Skeletal muscle Gastrocnemius – transverse section Peripheral nerves Sciatic nerve (proximal) – longitudinal and transverse sections Tibial nerve (proximal at knee) – longitudinal and transverse sections Tibial nerve (distal) and calf muscle branches – longitudinal and transverse sections

20.5.3  Schedule-Controlled Operant Behavior This guideline is designed to detect functional neurotoxic effects and is used on a case-by-case basis only for substances that have been shown to produce neurotoxic signs in other toxicity studies or are structurally related to neurotoxicants that affect performance, learning, or memory. The guideline gives basic guidance (Table 20.9); however, most technical details are not constrained to allow a flexible approach. The study may be acute (single dose) or repeated dosing (subchronic or chronic; i.e., up to 2 years in duration), and any appropriate route of exposure is acceptable. Although the guideline discusses fixedratio and fixed-interval responding as particular methods, it also states that “additional tests may be necessary to completely assess the effects of any substance on learning, memory, or behavioral performance.” In reality, should this

type of testing be required, the regulatory agency and registrant will be involved in detailed discussions about why the study is necessary/appropriate and also about the most appropriate study design to further investigate effects seen. The guideline provides a forum for discussion and not a detailed protocol. SCOB involves the maintenance of a behavior (typically lever pressing) by positive or negative reinforcement. Different rates and patterns of responding are controlled by the relationship between response and subsequent reinforcement. Fixed-ratio schedules are those in which a response is reinforced only after a specified number of responses. Fixed-interval schedules are those in which the first response is rewarded only after a specified amount of time has elapsed. To investigate effects on memory, experimental animals are trained to perform under a schedule of reinforcement and measurements of their operant behavior are made. The test substance is then administered, and

Chapter | 20  Regulatory Aspects of Acute Neurotoxicity Assessment

measurements of the operant behavior are repeated. For use of this test to study learning, animals may be trained following exposure. The studies require significant investment in terms of equipment and time because significant training (weeks to months) is usually required to obtain terminal baseline of responding. Personnel expert in the conduct, design, and interpretation of studies are also required and are not commonly available at contract research organizations. The method commonly uses food or water deprivation to motivate responding and requires prominent sensory and motor components, and all these aspects must be considered in the interpretation of the data. The methodology has been used to investigate cognitive effects of a wide range of agents, including metals, solvents, and pesticides (Anger et al., 1979; Christoph et al., 2000; Cory-Slechta et al., 1981; Moser and MacPhail, 1986).

20.5.4  Peripheral Nerve Function Nerve conduction studies can be useful in investigating possible peripheral neuropathy and are the most commonly used test of electrophysiological function in neurotoxicology (Johnson, 1980). This guideline is considered to be a “second tier” study for substances that have been shown to produce peripheral neuropathy (or other neuropathological change in peripheral nerves) in other neurotoxicity studies or compounds that show structural similarity to agents known to cause peripheral neuropathy. The purpose of the study is to record amplitude and velocity of conduction in peripheral nerves in vivo. Protocol details are similar to those for the neurotoxicity screening battery (Table 20.10) to allow a combination of the two components into a single study design. No specific detail is given on the timing of measurements, and these can be performed repeatedly in the same animal. However, if surgical exposure of the nerves is used, it is assumed that this is done at the end of the exposure and/or observation period. Nerve conduction velocity must be assessed in both sensory and motor nerve axons separately. Most peripheral nerves contain mixtures of individual sensory and motor nerve fibers. To distinguish sensory from motor effects, stimulating and recording electrodes are differentially positioned. For motor nerve conduction velocity (MCV), the stimulating electrodes are placed proximally near the nerve and the recording electrodes are placed near the muscle (to record the muscle response). For sensory nerve conduction velocity, both the stimulating and the recording electrodes are based close to the nerve, with the stimulating electrode distal to recording electrodes. Either a hindlimb (e.g., tibial) or tail (e.g., ventral caudal) nerve must be chosen for the guideline study. The other critical variable measured is amplitude of the evoked potential, which may be recorded from a mixed nerve. A difficulty in determining conduction velocities in intact animals is the accurate measurement

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Table 20.10  OPPTS 870-6850 Peripheral Nerve Function Species and age

Young adult rat 42–120 days old. Other species may be used if more appropriate.

Group size

At least 10 animals of one sex per group.

Control groups

Concurrent vehicle and/or untreated control required.

Positive control

Historic control data acceptable. Data should demonstrate sensitivity of methods and ability of methods to detect changes in peripheral nerve function.

Dose levels

At least three dose levels plus control. Highest dose need not exceed 2 g/kg.

Route of exposure

Most appropriate based on likely human exposure, bioavailability, and practicality. Typically gavage.

Observations

Measurement of motor and sensory nerve conduction velocity and response amplitude in vivo. Either a hindlimb (e.g., tibial) or tail (e.g., ventral caudal) nerve must be used. Core and nerve temperature, body weight.

Frequency of observations

Not specified.

of the conducting distance. In the rat, tail distances can be measured with sufficient accuracy. For other nerves such as the tibial, the uncertainty of electrode location introduces error into the measurements unless the nerve is surgically exposed. Normal peak conduction velocity measurements are influenced by a number of factors, the most important of which is temperature. It is therefore important to control body and nerve temperatures within a narrow range (0.5°C), and these parameters need to be recorded and reported. Decreases in peripheral nerve conduction velocity may be indicative of demyelination. For example, reduced MCV measurements in rats, guinea pigs, cats, and monkeys have been shown to be associated with segmental demyelination following acrylamide ­ exposure (Fullerton and Barnes, 1966; Leswing and Ribelin, 1969). However, because the large myelinated fibers conduct EPs faster than do smaller fibers, the maximum MCV provides a measure of the integrity of the larger, faster conducting fibers. Effects on slower conducting axons can be difficult to detect because MCV may be normal but action potential ­ amplitude

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may be decreased (De Jesus et al., 1978). Recording of response amplitude requires careful experimental technique (e.g., electrode placement), a larger sample size, and greater statistical power than measurements of velocity (WHO, 2001). Thus, the interpretation of MCV studies is enhanced if nerve pathology is also included. Decreases in response amplitude reflect a loss of nerve fibers and may occur prior to decreases in conduction velocity (De Jesus et al., 1978).

20.5.5  Neurophysiology: Sensory Evoked Potentials This guideline involves neurophysiologic measurement in vivo in adult rats to assess sensory function (Table 20.11). Such studies are unlikely to be required on a routine basis but may be used at any time to detect sensory dysfunction. Alternatively, the study may be requested on a case-bycase basis to clarify effects seen in other studies. Similar to the guideline for peripheral nerve function, the design is

Table 20.11  OPPTS 870.6855 Neurophysiology: Sensory Evoked Potentials Species and age

Young adult rat 42–120 days old, preferably pigmented. Implantation of chronic electrodes not recommended prior to 60 days. Other species may be used if more appropriate.

Group size

At least 10 animals of one sex per group.

Control groups

Concurrent vehicle and/or untreated control required.

Positive control

Historic control data acceptable. Data should demonstrate sensitivity of methods and ability of methods to demonstrate functional changes in the sensory systems to be tested.

Dose levels

At least three dose levels plus control. Highest dose need not exceed 2 g/kg.

Route of exposure

Most appropriate based on likely human exposure, bioavailability, and practicality.

Observations

Measurement of auditory, somatosensory, and/or visual evoked potentials in vivo. Body temperature at time of recording, body weight.

Frequency of observations

To include the estimated time of peak effect.

such that it may be used in combination with other studies (e.g., neurotoxicity screening battery or standard toxicity study) and may be of any duration (acute to chronic). The guideline details a number of stimuli that may be used to assess sensory function (visual, auditory, and somatosensory) and gives details of recording methods, but the specific details of study design are flexible to ensure that the most appropriate test methods are used. Indeed, the guideline states that it is “the responsibility of those submitting to justify the selection of a specific test from the categories of electrophysiological tests available.” SSEPs are elicited by electrical stimulation of sensory receptors or peripheral nerves at the foot, tail, or skin and are recorded from the somatosensory cortex. Recording from the cerebellum can help to differentiate effects and/or more precisely localize lesions. The SSEP records activity from the entire sensory pathway from the limbs to the brain (Mattsson and Albee, 1988; OTA, 1990). Early components are thought to represent the far-field dorsal column and thalamic activity and later components are of near-field cortical origin. VEPs [which include flash evoked potentials (FEPs) and pattern reversal evoked potentials) are used to evaluate effects on the parts of the nervous system responsible for vision. Potentials can be generated using stimuli ranging from diffuse light flashes to complex patterns of shape or color. If abnormalities are observed in FEPs, electroretinography, recorded using corneal electrodes, may be used to aid interpretation (Rebert, 1983). Auditory evoked potentials may be recorded from the cortex or the brainstem (BAER) in response to clicks and can be used to detect specific losses in the auditory system. BAERs generated using tone pips of varying frequencies and intensities are used to assess peripheral auditory dysfunction (Rebert, 1983). SEPs have a number of advantages for use in laboratory animals, including a fixed temporal relationship with the evoking sensory stimulus making them extremely reliable both within and between individuals. Because they reflect rather directly the integrity of sensory CNS pathways, they can yield specific information about particular neuroanatomical pathways and allow correlations of structure and function. They can be recorded noninvasively and thus can be recorded simultaneously with behavioral measures and repeatedly in the same individual. SEPs, like other evoked potentials, are not without interpretational difficulties because they may be altered by a variety of factors, such as toxic reactions, temperature, hypoxia, sensory deficits, central dysfunction, vitamin deficiency, or state of surroundings of nerve tract (Albee et al., 1987; Rebert, 1983; Sohmer, 1991). Therefore, it may sometimes be difficult to discriminate a direct neurotoxic effect from other consequences of treatment. Practical considerations in conducting the test include the specialist equipment and personnel required to conduct

Chapter | 20  Regulatory Aspects of Acute Neurotoxicity Assessment

the test and the need for surgery to implant electrodes. These factors again mean that the techniques are not commonly available at contract research organizations. Nevertheless, evoked potentials have been included in a large number of neurotoxicity studies (ECETOC, 1992; OECD, 2004) and have been shown to be useful indicators of neurotoxicity in animals.

20.5.6  Acute Cholinesterase Comparative Sensitivity Studies These studies are not described in formal guidelines but were requested to be conducted to provide data to assist in the risk assessments for organophosphorus pesticides (U.S. EPA, 2001). Under the September 19, 1999, data call-in for adult and developmental neurotoxicity studies for these agents, registrants were required to assess the adequacy of postnatal dosing during a developmental neurotoxicity (DNT) study. In a DNT study, pup exposure to test substance during the lactation period may occur via three pathways – maternal transfer via milk, consumption of treated diet by pups, and/or direct dosing of pups – and can be difficult to quantify accurately. Proposed indirect measurements – measurement of milk content of test substance or measurement of a biomarker (e.g., ChE inhibition) – or the presence of clinical signs (e.g., decreased body weight) do not provide data that allow an assessment of the relative sensitivity of adults and children as required by FQPA. Consequently, studies of comparative sensitivity in pups and adults were requested where exposure data in the DNT study were insufficient to adequately estimate the dose to pups. The primary goal of acute comparative sensitivity studies is to determine both the no-observed-adverse-effect level/lowest-observed-adverse-effect level and some measure of the dose–effect curve, such as ED50 for ChE inhibition in each of the three compartments (plasma, red blood cells, and brain) in both young and adult animals (see Table 20.12 for outline design). Although the design is relatively simple, there are a number of practical considerations that complicate the study conduct. These include the need to have the maternal animal available for preweaning pups following dosing and the need to control for litter (e.g., by including all dose groups within a single litter or cross-fostering animals across litters). Consequently, the number of litters required is usually increased. If all dose groups are included within a litter, then there is potential for cross-contamination or effects on nurturing if the high dose is associated with toxicity (for practical advice, see International Life Sciences Institute, 2003). Direct dosing of preweaning rat pups demands extreme care. Depending on the age of pup chosen, it may be necessary to pool blood samples. In addition,

599

Table 20.12  Acute Cholinesterase Comparative Sensitivity Study Species and age

Rat. Preweaning pups: early to midlactation (day 11); late lactation (day 21). Young adults.

Group size

At least five animals per sex per group at each age.

Control groups

Concurrent vehicle and/or untreated control required.

Positive control

Not required.

Dose levels

At least three dose levels plus control for each age.

Route of exposure

Gavage.

Observations

Activity of cholinesterase in plasma, red blood cell, and brain. Survival, body weight, and ageappropriate clinical signs.

Frequency of observations

At time of peak effect (determined in preliminary study).

the time of peak effect may be different between sexes, age groups, and compartments, necessitating increased numbers of groups and complications in study scheduling. Because of the number of animals required (minimum of 20/sex/age group) and requirements to time termination closely and remove, prepare, and analyze samples as quickly as possible, the studies typically have to be split over a number of days. Specific methodology for the cholinesterase assay is as described previously, although modifications to the method have been recommended by the U.S. EPA taking account of concerns about AChE assays raised by Wilson et al. (1996). In well-conducted studies, it is possible to detect statistically significant differences on the order of 5–10% in brain, 10% in red blood cells, and 20% in plasma.

Conclusion Significant progress has been made in recent years in the area of animal neurotoxicology testing. A wide range of animal toxicity study designs have been developed, validated, and used extensively to address potential neurotoxicity following acute exposures. Procedures for both behavioral and neuropathological evaluation are routinely available and frequently used in regulatory toxicity studies for risk assessment purposes. Special tests to assess chemicalinduced changes in sensory, motor, and cognitive function have also been used.

600

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Chapter 21

Proteomics in Pesticide Toxicology Su-wei Qi and Qing X. Li University of Hawaii, Honolulu, Hawaii

Pesticides are the cornerstones of pest management, food security, and public health. The main issues concerning pesticides include their proper usage, regulation, properties, efficacies, adverse effects, mechanisms of toxicity, modes of action, and the discoveries of effective and safe pesticides. Pesticide toxicology is the study of the adverse effects of pesticides on living organisms. Enzymes and receptors (i.e., proteins) are the primary, if not sole, targets of pesticides designed to control insects, weeds, fungi, bacteria, etc. Proteins are also the major targets of pesticides causing adverse effects on nontarget species such as humans and wildlife. Proteomics studies, therefore, are expected to contribute significant insights into complex issues of pesticide toxicology. The proteins expressed from a genome can be referred to as the “proteome” (Wasinger, 1995; Wilkins et al., 1997). The proteome of a biological system (biosystem; e.g., organism, cell, and organelle) is the ensemble of proteins expressed under a specific physiological condition, whereas the genome refers solely to the complete set of genes. Gene splicing and post-translational modifications further complicate proteomes—that is, the number of proteins exceeds that of genes, especially in eukaryotes. Therefore, the proteome is larger than the genome. The genome is defined by the sequence of nucleotides. The proteome, however, is not limited to the sum of the sequences of the proteins present; it also includes the structures and functions of the proteins. Proteomics is the study of the proteome or the entirety of proteins, particularly structures and functions of proteins in a biosystem. Proteomics involves identification and profiling of all proteins or as many as possible in a sample, quantitation of proteins, protein networks (e.g., interactions among proteins and between proteins and other biomolecules), protein folding, and protein modifications. Proteomics also involves expression of proteins at a particular physiological state of the organism or cell or in response to genetic manipulations. The number of proteomics publications has increased exponentially during Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

the past decade (Figure 21.1; see also Figure 22.2 in Chapter 22). Proteomics techniques have been applied to mechanism of toxicity, mode of action, identification of protein biomarkers of exposure, exposure monitoring, metabolisms of pesticides in humans and animals, and biodegradation of pesticides (Barrier and Mirkes, 2005; John, 2008; Nesatyy and Suter, 2007; Park and Lee, 2007). Proteomics applied to toxicology and pathology is referred to as toxicoproteomics. This chapter is an introduction of concept, method, and application of proteomics in pesticide toxicology, with specific examples of proteomics studies in bacterial degradation of pesticides.

21.2  Proteomics Methods Proteomics not only characterizes the final gene pro­ducts in a biosystem but also provides detailed information about protein abundances, stabilities, turnover rates, functions, structures, post-translational modifications, and protein– 450 400 Number of publications

21.1  Introduction

350 300 250 200 150 100 50 0

99 000 001 002 003 004 005 006 007 008 2 2 2 2 2 2 2 2 2

19

Year (1999–2008) Figure 21.1  Publications on pesticide proteomics in the past decade. The number of publications obtained with the key word of “pesticide proteomics” in the NCBI database has increased exponentially from 1999 to 2008.

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protein interactions. Large-scale, high-throughput omics technologies can comprehensively reveal complex protein networks in a biosystem (Hendrickson et al., 2008). Proteomics is associated with a suite of methods ranging from bioassays, molecular biological assays, genetic manipulations, bioreactor and bioengineering to nuclear magnetic resonance (NMR) and mass spectrometry (MS) analyses (Chen, 2008; Nesatyy and Suter, 2007; Shin et al., 2008). Proteomics studies require well-thought-out experimental designs, representative sampling, and proper sample handling and analyses (see Chapter 22). In addition to metabolomics, genomic techniques such as gene disruption, silencing, cloning, manipulation, amplification, and expression are complementary to proteomics (Campbell and Heyer, 2006; see Chapter 22). Among various protein analytical techniques, MS has emerged as the primary method (Cravatt et al., 2007; Siuti and Kelleher, 2007). MS-based proteomics can be classified as (1) top-down

(analysis of intact proteins) and (2) bottom-up (analysis of peptides after protein digestions) (Figure 21.2). Proteins or digested peptides can be separated with gel electrophoresis (gel-based) or liquid chromatography (gel-free) and subsequently identified and characterized by MS analyses.

21.2.1  Two-Dimensional Gel Electrophoresis Since the development of two-dimensional gel electrophoresis (2D-GE; O’Farrell, 1975), several technical innovations have improved the robustness and reliability of the technique (Lilley et al., 2002). Key advancements include the development of immobilized pH gradients (IPGs) (Celis and Gromov, 1999; Görg et al., 1988, 2000), sensitive staining methods (Rabilloud et al., 1992), and powerful image analysis systems (Westergren-Thorsson et al., 2001). IPGs have significantly improved the resolution and

Fractionation (e.g., FPLC; FFE)

Extraction Sample (e.g., bacteria)

Harvest Wash Precipitation Isolation, etc

Centrifugation

Cells Proteins

Extraction

Bottom-up

Top-down Gel-base

Supernatant

DNA/RNA

MudPIT

Gel-free

Separation Digestion

Extraction Filtration

Acidic/Neutral Derivaization

GC/MS

LC/MS (e.g., QQQ, Q-TOF)

DGGE PCR RT-PCR Sequencing Cloning Expression

2D-GE

DIGE

Image Analysis

MALDI TOF FTMS Orbitrap MS

1D-GE

Ion exchange column

Digestion

Digestion MALDI TOF

LC/IT MS

LC/IT MS

Databases Search and Amino Acid Sequence

Figure 21.2  Schematic of “multi-omics” approaches applied in metabolism studies of pesticides, where MS-based proteomics and metabolomics workflow is emphasized. Metabolites in the supernatant of bacterial sample can be analyzed by GC/MS and LC/MS. DNA and RNA extracted from cells can be analyzed by DGGE, PCR, RT-PCR, sequencing, cloning, and other molecular techniques. Proteins extracted from bacterial cells can be analyzed via top-down or bottom-up approaches. In the bottom-up approach, a mixture of proteins can be separated on SDS–PAGE and then analyzed by MS. Alternatively, the protein mixture can be directly digested into a collection of peptides that are then separated and determined by multidimensional chromatography on-line coupled to tandem mass spectrometric analyses [i.e., multidimensional protein identification technology (MudPIT)]. In the top-down approach, intact proteins are fractionated into less complex protein mixtures for MS analysis. Bioinformatics including database search and amino acid sequence alignment are then conducted to identify and characterize the proteins and peptides. DGGE, denatured gradient gel electrophoresis; DIGE, difference gel electrophoresis; FFE, free flow electrophoresis; FPLC, fast protein liquid chromatography; FTMS, Fourier transform mass spectrometry; LC/MS, liquid chromatography/mass spectrometry; MALDI TOF, matrix-assisted laser desorption ionization time-of-flight; QQQ, triple quadrupole; Q-TOF, quadrupole-time-of-flight; RT-PCR, real-time polymerase chain reaction.

Chapter | 21  Proteomics in Pesticide Toxicology

reproducibility. The integration of high-resolution 2D-GE, MS detection, and bioinformatic data processing is a wellestablished proteomics workflow (Aebersold and Mann, 2003; Agrawal and Thelen, 2009; Chalkley et al., 2005; Dunn and Görg, 2001; Stults and Arnott, 2005). A typical 2D-GE–MS workflow consists of (1) protein extraction and fractionation, (2) separation of proteins on 2D-GE, (3) staining and spot detection, (4) image analysis, (5) MS protein profiling, and (6) protein database search based on MS and 2D-GE data (see Figure 21.2). Resolving a large number of proteins requires close technical attention to each step in 2D-GE. Extraction of proteins must be efficient, quantitative, and reproducible for all (low and high abundance) proteins or a particular type of protein. Table 21.1 shows essential reagents for unfolding and denaturation of proteins. Protein degradation (e.g., hydrolysis and oxidation) and modification (e.g., carbamylation) must be minimized prior to 2D-GE. Membrane proteins are one of the most difficult protein classes because of their hydrophobicity and embedment in the lipid bilayers (Santoni et al., 2000). Rabilloud et al. (2008) reviewed the applications of 2D-GE and zone electrophoresis for separation of membrane proteins. Cathode drifting and pH flattening near the anode of highly basic and acidic proteins remain challenges in 2D-GE runs, although pH-segmented IPGs improve the resolution of proteins. Table 21.2 shows common protein staining methods. Most of the staining methods are compatible with MS analysis and afford good reproducibility. Limits of detection and linear ranges vary among staining methods (Walker, 2002).

605

Proteins can be labeled with cyanine fluorescence dyes (e.g., Cy2, Cy3, and Cy5) prior to 2D-GE (Unlü et al., 1997). Differential in gel electrophoresis (DIGE) has improved the reproducibility of 2D-GE by minimizing gel-to-gel variations (Alban et al., 2003; van den Bergh and Arckens, 2004). Cyanine dyes are individually mixed with different protein samples in vitro prior to isoelectrofocusing. Fluorescent cyanine dyes that differ in their excitation and emission wavelengths allow image overlay and normalization for qualitative and quantitative analyses of 2-D gels. DIGE and traditional 2D-GE are most commonly combined with matrix-assisted laser desorption ionization time-of-flight MS (MALDI TOF).

21.2.2  Chromatographic Separation of Proteins Chromatographic separation of proteins is an alternative to electrophoresis. Figure 21.2 shows some common MSbased proteomics methods. Shotgun proteomics is a method of identifying proteins in complex mixtures using a combination of liquid chromatography (LC) and MS, providing a wider dynamic range and coverage of proteins than 2D-GE, particularly for hydrophobic proteins and low-copy proteins (Hendrickson et al., 2008; Lee and Lee, 2004; Shimizu, 2004). Shotgun approaches, therefore, are widely used to catalogue proteomes and characterize post-translationally modified proteins (Denny et al., 2008; Kang et al., 2008). A disadvantage of the shotgun approach is its difficulty in protein quantitation. Semiquantitative analyses of proteins can be

Table 21.1  Essential Reagents for Unfolding and Denaturation of Proteins Prior to 2D-GE Purpose

Reagents

Suggested concentration

To disrupt hydrogen bonds

Urea Thiourea

8–9.8 M 2M

To inhibit protease To break intra- and intermolecular disulfide bonds

Protease inhibitor DTE/DTT Iodoacetamide Tributylphosphine

40 mM 2% w/v 2 mM

To block hydrophobic interaction

SDS 16-BAC CHAPS ASB-14 ASB-16 ASB-C80 Zwittergent l--Lysophosphatidylcholine decanoylN-methylglucamide Nonidet P-40

20% 7.5% w/v 4% 2% 2% 1% 1% 0.5–2%

ASB, amidosulfobetaine; CHAPS, 3-[(3-cholamidopropyl)dimethylammonio]-1-propanesulfonate; DTE, dithioerythritol; DTT, dithiothreitol; SDS, sodium dodecyl sulfate; 16-BAC, benzyldimethyl-n-hexadecylammonium chloride.

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Table 21.2  Common Protein Staining Methods Used in 2D-GE Staining method

Dye

Postelectrophoretic stains Coomassie CBB-R bright blue CBB-G (colloidal) Silver stain Silver nitrate (acidic) Silver ammonia (alkaline) Negative staining Zinc/imidazole Fluorescence SYPRO Ruby SYPRO Orange, Red, and Tangerine Epicoccone (Lightning Fast, Deep Purple) Preelectrophoretic stains Fluorescence CyDyes minimal labeling CyDyes saturation labeling FlaSHPro Dyes

LOD (ng)

Linearity (orders of magnitude)

MS compatibility

8–100 10 1–5 1 1–10 1 4–10

1–1.3



1–2 1–2

a 

3 3

1

4

0.1–2 0.005–0.01 2–3

3–5 3–5

 



  

CBB-G, Coomassie bright blue G250; CBB-R, Coomassie bright blue R250; LOD, limit of detection. a Silver stain is compatible with MS when no glutaraldehyde is added.

achieved with multidimensional protein identification technology (MudPIT; Chen and Harmon, 2006). The protein mixture is directly digested into peptides that are then resolved by multidimensional chromatography on-line coupled to tandem MS (MS-MS) (i.e., MudPIT). Peptide mixtures are concentrated and cleaned up on an ion-exchange column, separated on a reverse-phase column, and then subjected to MS-MS (MacCoss et al., 2002; Washburn et al., 2001). 1D-GE coupled with LC is referred to as GeLC (Kristjansdottir et al., 2008; Schirle et al., 2003). A complex protein mixture is separated on 1D-GE, visualized after staining, sliced, and then subjected to in-gel enzymatic digestion such as trypsin. Tryptic peptides are extracted and analyzed on reverse-phase nano-LC/MS and MS-MS (Myint et al., 2009; Yang et al., 2007). Because GeLC is a combination of two separation mechanisms, it is superior to MudPIT and other techniques with regard to the number of proteins identified (Breci et al., 2005; Yates et al., 2009). Table 21.3 lists common software and websites for the analysis of 2D-GE and LC/MS images.

21.2.3  Stable Isotope Labeling of Proteins Protein quantitation is an active proteomics research topic. Tagging of stable isotope probes onto proteins and peptides facilitates their quantitative measurement. Gygi et al. (1999) reported the isotope-coded affinity tag (ICAT) labeling technique for quantitative analyses of protein mixtures. The ICAT reagent contains a biotin (for interaction with

Table 21.3  Common Software and Websites for the Analysis of 2D-GE Images and LC/MS Images Software

Website

2D-GE image analysis DeCydera Delta2Da ImageMaster 2D Platinuma PDQuesta Progenesisa (formerly Phoretix) Proteomweavera Flickerb

http://www.gehealthcare.com http://www.decodon.com http://www.gehealthcare.com http://www.bio-rad.com http://www.nonlinear.com

http://www.definiens.com http://www.lecb.ncifcrf.gov/flicker/wgFlkPair. html GelScapeb http://www.gelscape.ualberta.ca ImageMaster 2D http://www.expasy.org Platinum and Melanie Viewerb

LC/MS image analysis DeCyder MSa MapQuantb Msightc MsInspectb Mzmineb OpenMSb SpecArrayc XCMSb a

http://www.gehealthcare.com http://arep.med.harvard.edu/MapQuant http://www.expasy.org/MSight https://proteomics.fhcrc.org/CPAS http://mzmine.sourceforge.net http://open-ms.sourceforge.net http://sourceforge.net/projects/sashimi/files/ SpecArray/SpecArray%20v1.1 http://metlin.scripps.edu/download

Commercialized product. Open-source package. c Free software. b

Chapter | 21  Proteomics in Pesticide Toxicology

avidin), an iodoacetamide group (coupling with reduced cysteine residues), and a spacer labeled with either light (e.g., hydrogen) or heavy (e.g., deuterium) isotopic variants (Barrier and Mirkes, 2005; Hägglund et al., 2008; Haqqani et al., 2008). Hansen et al. (2003) developed a cleavable ICAT (cICAT) reagent that allows for the release of the proteins from biotin–avidin immobilized columns. ICAT and cICAT methods exclusively label cysteine residues; therefore, cysteine-free proteins that account for approximately 20% of the proteome are excluded from the analysis (Schmidt et al., 2004). Isobaric tags for relative and absolute quantitation (iTRAQ) allow for amine-specific isobaric tagging and thus are suitable for simultaneous analysis of proteins regardless of the presence of cysteine residues (Boehm et al., 2007; Hundertmark et al., 2009; Ross et al., 2004). Chemical tagging methods such as ICAT, cICAT, and iTRAQ are in vitro labeling techniques that require subsequent enrichment, purification, and MS analyses of the proteins. A novel metabolic labeling method, stable isotope labeling by essential amino acid culture (SILAC), has emerged as an alternative by which arginine and lysine are labeled in vivo (Ong and Mann, 2006; Ong et al., 2002). SILAC accurately probes signal transduction networks and discerns true protein interactions (Guo et al., 2008; Krüger et al., 2008; Vermeulen et al., 2007). Software packages are available to calculate relative ratios of proteins and peptide pairs labeled with these tagging methods (Table 21.4). ASAPRatio (Li et al., 2003), XPRESS (Han et al., 2001), and RelEx (MacCoss et al., 2003) calculate the relative abundances of ICAT-labeled peptides after their analysis with LC/MS-MS and identification with SEQUEST or MASCOT. The MSQuant quantifies SILAC-labeled peptides, and ZoomQuant is specialized in the quantitation of 18 O-labeled peptides after identification with SEQUEST (Halligan et al., 2005).

21.2.4  Mass Spectrometry Identification of Proteins A key value of MS-based proteomics is its accurate identification of proteins of interest. Electrospray ionization (ESI) and MALDI are key ionization methods for protein and peptide analysis. Mass analyzers used in proteomics are TOF, quadrupole (Q), ion trap (IT), Fourier-transform ion cyclotron resonance (FT-ICR), Orbitrap, as well as their combinations. MS-MS fragmentation methods include collision-induced or collisionally activated dissociation, electron capture dissociation, and electron transfer dissociation. Chen (2008) provides an excellent review of MS-based proteomics. After MS detection, proteins or peptides can be identified via peptide mass fingerprinting (PMF), peptide fragmentation fingerprinting (PFF), and de novo sequencing

607

(Table 21.4). In PMF analysis, the experimental mass spectra from an enzymatically digested peptide are compared with theoretical data computed from the amino acid sequences available in databases and in silico digested using the same cleavage specificity of the protease employed in the experiment. MOWSE (molecular weight search), for example, compares the calculated peptide masses for each entry in the sequence database with the experimental data and uses empirically determined factors to assign a statistical weight to each individual peptide match (Pappin et al., 1993), which is exploited in MS-Fit (Clauser et al., 1999) and MASCOT (Perkins et al., 1999). Comprehensive reviews of these and other related scoring functions are available (Gras and Muller, 2001; Hernandez et al., 2006). Any given peptide ion can be selected and further fragmented on MS-MS, giving MS-MS spectra, also called PFF. The PFF approach compares MS-MS peptide spectra with theoretical spectra in databases. Some algorithms have been developed to reduce the number and complexity of MS-MS spectra while increasing their quality [e.g., NoDupe (Tabb et al., 2003)]; to handle unexpected PTM or mutations [e.g., Popitam (Hernandez et al., 2003, 2006), GutenTag (Tabb et al., 2003), and InsPecT (Tanner et al., 2005)]; and to deal with special modifications using various strategies [e.g., Phenyx (Palagi et al., 2009) and X!Tandem (Craig and Beavis, 2004)]. Shadforth et al. (2005, 2006) reported a value-based scoring system that evaluates PMF-based protein identification without accompanying amino acid sequence data from MS-MS analysis. De novo peptide sequencing is defined as peptide sequencing performed without prior knowledge of the amino acid sequence. Table 21.4 presents a number of algorithms and tools dedicated to de novo sequencing. DeNovoX and Spectrum Mill are commercial software installed in the MS equipment. PEAKS is a stand-alone commercialized software; it also has a web interface for free submissions. De novo sequencing is mainly used for cross-species identification (Liska and Shevchenko, 2003) or in the PFF approach to generate partial sequence information to filter candidate peptides prior to identification (Frank et al., 2005). It is noteworthy that manual reviews and verification of the automatic research results are required. Numerous proteomic data processing platforms have been implemented to automate the identification process and, thus, to reduce data analysis time, enhance the quality of identification, and increase the coverage of matched spectra (Palagi et al., 2006).

21.2.5  Protein Bioinformatics Protein bioinformatics includes experimental and computational approaches ranging from simple comparison of protein amino acid compositions to sophisticated software

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Table 21.4  Software and Websites for MS-Based Protein Identification and Quantitation Software

Website

PMF tools Aldente MASCOT MS-Fit PepFrag PepMAPPER PeptideSearch ProFound

http://www.expasy.org/tools/aldente http://www.matrixscience.com http://prospector.ucsf.edu http://prowl.rockefeller.edu/prowl/pepfrag.html http://wolf.bms.umist.ac.uk/mapper http://www.unb.br/cbsp/paginiciais/peptsrcseq.htm http://prowl.rockefeller.edu/prowl-cgi/profound.exe

PFF tools GutenTag InsPecT MASCOT MS-Tag and MS-Seq NoDupe OMSSA PepFrag PepProbe Phenyx Popitam ProID SEQUEST Sonar MS-MS Spectrum Mill VEMS X!Tandem

http://fields.scripps.edu/GutenTag/index.html http://proteomics.ucsd.edu/InspectDocs http://www.matrixscience.com/search_form_select.html http://prospector.ucsf.edu http://fields.scripps.edu/nodupe/index.html http://pubchem.ncbi.nlm.nih.gov/omssa http://prowl.rockefeller.edu/prowl/pepfrag.html http://bart.scripps.edu/public/search/pep_probe/search.jsp http://www.genebio.com/products/phenyx http://www.expasy.org/tools/popitam http://sashimi.sourceforge.net/software_mi.html http://fields.scripps.edu/sequest/index.html http://hs2.proteome.ca/prowl/sonar/sonar_cntrl.html http://www.home.agilent.com http://yass.sdu.dk http://human.thegpm.org/tandem/thegpm_tandem.html

MS-MS de novo sequencing tools AUDENS DeNovoX Lutefisk PEAKS PepNovo Sequit! Spectrum Mill

http://www.ti.inf.ethz.ch/pw/software/audens/ http://www.thermo.com http://www.hairyfatguy.com/lutefisk http://www.bioinformaticssolutions.com http://proteomics.ucsd.edu/Software/PepNovo.html http://www.sequit.org http://www.home.agilent.com

Pipeline tools ProteinLynx Global Server ProteinScape Scaffold Trans-Proteomic Pipeline

http://www.waters.com/waters/nav.htm?localeen_US&cid10053564 http://www.proteinscape.com http://www.proteomesoftware.com http://tools.proteomecenter.org/wiki/index.php?titleSoftware:TPP

Tools for MS data analysis ASAPRatioa DTASelectb MSQuanta PeptideProphetb ProteinProphetb RelExa XPRESSa ZoomQuanta

http://tools.proteomecenter.org/wiki/index.php?titleSoftware:ASAPRatio http://fields.scripps.edu/DTASelect http://Msquant.sourceforge.net http://tools.proteomecenter.org/wiki/index.php?titleSoftware:PeptideProphet http://tools.proteomecenter.org/wiki/index.php?titleSoftware:ProteinProphet http://fields.scripps.edu/relex http://tools.proteomecenter.org/wiki/index.php?titleSoftware:XPRESS http://proteomics.mcw.edu

a

This software is particularly useful for quantitative analysis. This software is particularly useful for validation of protein identifications.

b

for large-scale protein profiling and structure elucidation. Bioinformatics tools include software for 2D-GE analysis and both qualitative and quantitative analyses of MS protein data (see Tables 21.3–21.5). Bioinformatics enables automated processes of MS analysis and enhances the quality

of the results (Palagi et al., 2006). During the past three decades, software packages have been developed to analyze 2D-GE images for: (1) detection and semiquantification of protein spots on 2D gels; (2) localization of protein spots within a gel; (3) matching of corresponding spots

Chapter | 21  Proteomics in Pesticide Toxicology

Table 21.5  Major Proteomic Databases Database website

Sourcesa

http://www.expasy.ch/ch2d/

Swiss 2DPAGE

http://www-lecb.ncifcrf. gov/2dwgDB

2DWG Image Meta-Database

http://bioinformatics.icmb. utexas.edu/OPD

Open Proteomics Database

http://www.systemsbiology.org

Institute for Systems Biology

http://www.sbeams.org

SBEAMS

http://www.expasy.ch/ch2d/ 2d-index.html

Index

http://mips.gsf.de

MIPS

http://www.bind.ca

BOND

http://dip.doe-mbi.ucla.edu

Database of Interacting Proteins (DIP)

http://www.ebi.ac.uk/intact

PPI database IntAct

http://www.mysql.com

MySQL

http://ca.expasy.org/tools/ pi_tool.html

Compute pI/Mw tool

http://spock.jouy.inra. fr/RL000801.html

MOLOKO

http://www.ncbi.nlm.nih.gov/ entrez/query.fcgi?dbProtein

NCBI protein database

http://www.wzw.tum.de/ proteomik/lactis

DynaProt 2D

http://www.cebitec. uni-bielefeld.de/cebitec/ computational-genomics/ software.html

BRIGEP

http://compbio.mcs.anl. gov/sentra/

SENTRA (Signal transduction proteins)

http://www.boutell.com/gd

GD library for 2D gels

609

new bioinformatics tools for simultaneous processing of MS and 2D-GE data as well as other protein information. Many protein, proteomics, and nucleotide databases are available on the World Wide Web (Kremer et al., 2005; see http://ca.expasy.org). Proteomics databases and database updates can be found in the 2009 database issue of Nucleic Acids Research (http://nar.oxfordjournals.org). Typical 2DGE databases consist of gel images obtained from specific organisms and allow users to select protein spots for more information to match and identify proteins of interest. Most of these databases also provide “clickable” map functionality (e.g., http://ca.expasy.org) based on the protein information, gel data, and spot coordination. The 2D library can dynamically display 2D-GE spots on the reference gels (Kremer et al., 2005). DynaProt 2D is another advanced proteomic database for dynamic online access to proteomes and 2D gels based on spot identification and annotation (Drews and Görg, 2005). In addition to experimental proteomic data, complete theoretical proteomes can be retrieved from databases such as the PAD (Proteome Analysis Database) or generated by calculating theoretical proteome maps (Pruess et al., 2003). The BRIGEP bioinformatics software system consists of three web-based applications: GenDB, EMMA, and ProDB (Goesmann et al., 2005). These applications facilitate the processing and analysis of bacterial genome, transcriptome, and proteome data. The SPD (Secreted Protein Database) covers a collection of secreted proteins from human, mouse, and rat proteomes, including sequences from SwissProt, TrEMBL, Ensembl, and Refseq (Chen et al., 2005). The CEBS (Chemical Effects in Biological Systems) is an integrated public repository for toxicogenomics data, including the study design and timeline, clinical chemistry and histopathology findings, and microarray and proteomics data (Waters et al., 2008). The PRIDE (Proteomics Identifications Database; http://www.ebi.ac.uk/ pride) was first described in 2006 and has been linked with data from the Human Proteome Organization projects as well as iTRAQ quantitative data (Jones et al., 2008).

a

Revised based on Singh (2006).

between gels; and (4) differential comparison of protein expression. Packages were available to the public in the early 1980s, and some of them have survived over the course of the computational evolution. Among these are PDQuest (based on Quest) (Garrels et al., 1989) and ImageMaster 2D Platinum (based on Melanie) (Appel et al., 1991) (see Table 21.3). Software programs such as DeCyder (Anderson et al., 2007; Krogh et al., 2007) and Progenesis SameSpots (Kang et al., 2009) are specialized for 2D-DIGE images. LC/MS data in two dimensions [i.e., the retention time and mass-to-charge ratio (m/z)] can be translated into an image to visualize differences in protein expression and to discover new proteins (Berger et al., 2002; Palmblad et al., 2002). MS workflows and 2D-GE images are complementary; thus, it is pertinent to develop

21.3  Applications of proteomics in pesticide studies Proteomics has been widely used to elucidate various interactions between pesticides and their target proteins in plants, fungi, insects, and animals. The number of publications on pesticide proteomics has increased dramatically during the past decade (see Figure 21.1). Many such studies have been conducted with well-known model organisms such as Arabidopsis thaliana, Caenorhabditis elegans, Drosophila melanogaster, Pseudomonas putida, Saccharomyces cerevisiae, and Triticum tauschii. There are several classes of herbicide-regulated proteins, most of which are involved in general stress responses. The large subunit of the ribulose1,5-bisphosphate carboxylase/oxygenase (RuBisCo) is significantly decreased in rice leaves after glyphosate treatment,

610

whereas glyphosate may cause oxidative stress as antioxidant enzymes including peroxidases and glutathione S-transferases (GSTs) are up-expressed (Ahsan et al., 2008). Several photosynthesis-related proteins, including several fragments of RuBisCo, displayed differential expression in grapevine treated with the herbicide flumioxazin (Castro et al., 2005). Gershater et al. (2007) used proteomics approaches to successfully elucidate a herbicide bioactivation role of carboxylesterase (AtCXE12), hydrolyzing the pro-herbicide methyl-2,4-dichlorophenoxyacetate to the phytotoxic 2,4dichlorophenoxyacetic acid (2,4-D) in A. thaliana. Teixeira et al. (2005) suggested that a number of elements were involved in a set of responses to 2,4-D in yeast: (1) upregulation of the antioxidant enzyme Ahp1p and the heat shock proteins Hsp12p and Ssb2p (or Ssb1p) related to herbicide stress; (2) enzymes involved in protein (Cdc48p) and mRNA (Dcp1p) degradation; (3) alterations related to carbohydrate metabolism (Eno1p, Eno2p, and Glk1p); (4) the disturbance of vacuolar and plasma membrane via vacuolar H()-ATPase (V-ATPase) function (Vma1p and Vma2p); and (5) the increased expression of amino acid biosynthetic enzymes (Arg1p, Aro3p, Aro8p, Gdh1p, His4p, Ilv3p, and Met6p) correlated with the decrease in amino acid concentrations. Braconi et al. (2009) found protein carbonylation as a biomarker of 2,4-D stress in wine yeast. The induction of the pathogenesis-related proteins, cell division proteins, and redox-mediating proteins plays an essential role in cell defense mechanisms against the herbicide flumioxazin in grapevine and 2,4-D in legume (Castro et al., 2005; Holmes et al., 2006). It is noteworthy that 3,5-dichlorocatechol, a 2,4-D metabolite, is active as an uncoupler of oxidative phosphorylation that represses the synthesis of ferric uptake regulator (Fur)-dependent proteins (e.g., fumarase C and l-ornithine N5-oxygenase) involved in the oxidative stress response and iron uptake in P. putida KT2440 (Benndorf et al., 2006). Benndorf et al. (2007) developed a metaproteome protocol enabling the study of the functional diversity of environmental microbial communities. Herbicide safeners protect cereal crops from herbicide toxicity. Zhang et al. (2007) studied the responses of proteins to the safener cloquintocet-mexyl protecting seedlings from injury by the herbicide dimethenamid in the diploid wheat T. tauschii seedlings and identified 29 safener-induced and 10 herbicide-regulated proteins. Safener-responsive proteins, mostly involved in xenobiotic detoxification, also included several new proteins, whereas herbicide-regulated proteins were categorized into several classes involved in general stress responses. Riechers et al. (2003) found that safeners protected grass crops from herbicide injury by dramat­ ically inducing the expression of GSTs. Zhang and Riechers (2004) reported that the herbicide safener fluxofenim dramatically induced overproduction of 18 proteins, including GSTs and aldo/keto reductase homologs involved in glycolysis and the tricarboxylic acid (TCA) cycle in T. tauschii seedlings. One GST, AtGSTU19, was induced only by the herbicide safener benoxacor in Arabidopsis seedlings (Smith

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et al., 2004). Four GSTs (AtGSTF2, AtGSTF6, AtGSTF7, and AtGSTU19) were significantly more abundant in copper (oxidative stressor)-treated seedlings. GSTs are a class of important stress-responsive enzymes in animals and insects. Alias and Clark (2007) reported the GST proteomes of adult D. melanogaster in response to the herbicide paraquat and the anticonvulsant phenobarbital as chemical stimuli. Paraquat increased the total GST activity and the relative amounts of the DmGSTs D1, D3, E6, and E7 isoforms, whereas phenobarbital increased the relative amounts of the D1, D2, E3, E6, E7, and E9 isoforms with a possible depression in the relative amount of GSTS1. MSbased proteomics approaches have proven useful in identification of pesticide–protein adducts. Dooley et al. (2008) identified many protein adducts formed in the pituitary gland of atrazine-exposed rats and in diaminochlorotriazine (an atrazine metabolite)-exposed LbetaT2 rat pituitary cells, including protease activator complex subunit 1, ubiquitin carboxyl-terminal hydrolase isozyme L1, tropomyosin, ERp57, and RNA-binding proteins. It is interesting to note that all these proteins contain active-site or solvent-exposed cysteine residues, making them vulnerable to covalent adduct formation. Grossmann (2005) provides a comprehensive review of the physiological profiling (“physionomics”) of the complex plant system to screen new herbicides, with an unknown mode of action. Integration of physionomics with functional genomics, transcriptomics, proteomics, and metabolomics promises to enable efficient identification of the mode of action of a new herbicide and facilitate the discovery process. Mancozeb is a dithiocarbamate fungicide. A study that used the toxicoproteomics approach indicated that the target genes and the putative main transcription activators of the complex mancozeb-induced expression changes are related to yeast response to stresses, particularly to oxidative stress, protein translation initiation and protein folding, disassembling of protein aggregates, and degradation of damaged proteins in the yeast S. cerevisiae (Santos et al., 2009). 2D-GE hyphenated with MALDI TOF was used to detect differential expression of 21 proteins in DDTsusceptible and -resistant lines of D. melanogaster. These proteins are putatively involved in biochemical pathways such as glycolysis and gluconeogenesis, the pentose phosphate pathway, the TCA cycle, and fatty acid oxidation (Pedra et al., 2005). Changes in the respective metabolic pathways appear to be the mechanism of resistance in the DDT-resistant Drosophila. Proteomic approaches revealed that DDE induces tissue-specific oxidative stress in the clam Ruditapes decussatus, which includes protein carbonylation and heat shock response (Dowling et al., 2006). Entomopathogenic organisms produce hydrolases that degrade insect exoskeletons, making them a viable alternative to conventional insecticides. Murad et al. (2008) characterized a suite of proteins including proteases, reductases, and acetyltransferases secreted from Metarhizium anisopliae, an ascomycete that is parasitic to the insect pest Callosobruchus maculatus. Rotenone, a natural pesticide

Chapter | 21  Proteomics in Pesticide Toxicology

derived from the jicama vine plant, causes Parkinson’s disease-like symptoms in rats. Jin et al. (2007) used the SILAC technique combined with polyacrylamide gel electrophoresis (PAGE) and LC/MS-MS to identify the mitochondrial protein profiles of mouse dopaminergic embryonic stem cells exposed to rotenone. They identified 1722 proteins, 110 of which displayed significant changes in relative abundance after rotenone treatment. The secondary metabolites-mediated interactions between organisms are the subject of allelopathic studies. These metabolites are generally produced and exuded from plants into the environment and may confer a competitive advantage. A detailed discussion of achievements in genomics and proteomics in allelopathy can be found in Macías et al. (2007). The proteome is highly dynamic; it varies with the species and responds to environmental influences; thus, it allows accurate distinction between the parental and transgenic lines. Kubis et al. (2004) used a combination of genomics, transcriptomics, and proteomics methodologies to successfully uncover functional specialization among the Arabidopsis Toc159 family of chloroplast protein import receptors. Proteomics was used to identify unintended side effects occurring in transgenic maize seeds that were attributed to genetic modifications. A total of 43 proteins are differentially expressed in transgenic seeds with respect to the isogenic controls (T06 vs. WT06). The proteome changes can be linked directly to the insertion of a single gene (Zolla et al., 2008). Scossa et al. (2008) carried out a parallel transcriptional and proteomic comparison of seeds from a transformed bread wheat line that overexpressed a transgenic low-molecular-weight glutenin subunit gene relative to the corresponding isogenic wheat. Proteomic analyses showed differential accumulation of several classes of endosperm proteins in the transformed endosperm during seed development. The levels of the endogenous glutenins and all subclasses of gliadins are diminished during seed filling in the transgenic genotype caused by the upregulation of the transgene and subsequent overexpression of the corresponding protein subunit. Protein toxins produced by Bacillus thuringiensis (Bt) are very toxic to many insects and have been used to control insect pests since the 1920s (Sharma, 2008). Bt toxins are considered environmentally friendly and nontoxic to humans and wildlife, including beneficial arthropods (National Research Council, 2002). Bt crops (e.g., corn, cotton, and soybean) are widely planted (Lemaux, 2008). The controversy surrounding Bt crops concerns possible genetic contamination, hypothetical gene flow, possible impact on nontarget species, and pest resistance to Bt toxins (Babendreier et al., 2005; Rosi-Marshall et al., 2007; Rui et al., 2005; Serratos-Hernández et al., 2007; Tabashnik et al., 2008). Many proteomics studies have been carried out to address these issues (Crickmore, 2005; Jurat-Fuentes and Adang, 2007). Griffitts et al. (2001) discovered that loss of the gene (C. elegans bre-5) encoding a putative -1,3-galactosyltransferase conferred resistance to the Bt

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toxins Cry5B and Cry14A, signifying Bt toxin resistance relevant to loss of carbohydrate modification. Candas et al. (2003) utilized a proteomic approach to examine changes in the gut proteins from the larvae of an Indianmeal moth (Plodia interpunctella) colony exhibiting resistance to Bt. A number of detected changes in the expression of midgut proteins indicate increased glutathione utilization, elevation in oxidative metabolism, and differential maintenance of energy balance within the midgut epithelial cells of the Bt-resistant Indianmeal moth larva. The results indicate that variations in amino acid content or modifications of certain proteins such as F(1)F(0)-ATPase are also important mechanisms of resistance to Bt toxin in the Indianmeal moth. Another mechanism of Bt resistance is a dramatic down-expression of chymotrypsin-like proteinase in the midgut of the Bt-resistant Indianmeal moth larva, signifying a reduction of chymotrypsin activity and a subsequent decrease in activation of Bt toxin in the insect midgut.

21.4  Microbial degradation of pesticides Microorganisms have an exceptional ability to exploit inorganic or organic chemicals for their growth (Alexander, 1999; Klein, 2000; Timmis, 2009). A large number of microbes (e.g., bacteria and fungi) have been isolated and studied for biodegradation of environmental pollutants such as pesticides and polycyclic aromatic hydrocarbons (PAHs) under aerobic conditions (Seo et al., 2009; Timmis, 2009). Although anaerobic degradation of pesticides in sediments is significant for aquatic life, little attention has been paid to this area. Dehalococcoides species can undertake dehalorespiration of halogenated compounds such as perchloroethylene (Maymó-Gatell et al., 1997; Seshadri et al., 2005), chlorinated benzenes (Adrian et al., 2000), and polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDDs and PCDFs) (Bunge et al., 2003). Dehalobacter, Desulfitobacterium, and Sulfurospirillum species can also perform reductive dehalogenation (Holliger et al., 1998; Loffler et al., 1996; Luijten et al., 2003). Microbial reductive biodegradation of aromatic, halogenated aromatic, and nitroaromatic compounds has been well reviewed (Foght, 2008; Klein, 2000; Kulkarni and Chaudhari, 2007; Spain et al., 2000; Watrous et al., 2003). Advances in proteomics have allowed a comprehensive examination of abundance and global profile of proteins, adaptation, and the identification of key metabolic enzymes and regulatory proteins in microorganisms at a given physiological state. Pseudomonas putida KT2440, for example, has become a model bacterium for proteomic studies because of its metabolic versatility and the availability of its complete genome sequence (Nelson et al., 2002). Heim et al. (2003) constructed the first proteome reference map from P. putida KT2440 cultured in a mineral salt medium supplemented with glucose. In addition to

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transcriptome analyses (Miyakoshi et al., 2007) and functional genomics studies (Reva et al., 2006), differential proteomes were determined in P. putida KT2440 exposed to glucose, succinate, pyruvate, phenol (Kurbatov et al., 2006; Santos et al., 2004), and chlorophenoxy herbicides (Benndorf et al., 2006). Under aerobic conditions, key mechanisms of response to chemical stresses in P. putida include adaptation of the membrane barrier, uptake of phosphate, maintenance of the intracellular pH and redox status, oxidative stress response, energy metabolism, uncoupling of oxidative phosphorylation, inhibition of cell division, transport of small molecules, and regulation of translation and transcription. The following discussion focuses on proteomics in bacterial biodegradation of pesticides, regulation of metabolism, membrane proteins, and adaptation (Table 21.6).

21.4.1  Proteomics in Bacterial Degradation of Pesticides Bacterial degradation consists of many types of reactions, such as oxidation, hydrolysis, reduction, and dehalogenation

(Alexander, 1999). For example, alkanes undergo an initial monooxidation to become an alcohol and then carboxylic acid followed by -oxidation in microorganisms under aerobic conditions. Ring dioxygenation is a major initial reaction, catalyzed by aromatic ring hydroxylating (ARH) dioxygenases, on an aromatic ring in a bacterium under aerobic conditions. Biodegradation pathways that consist of many steps of reactions can be broadly divided into peripheral and central pathways (Alexander, 1999; van der Meer et al., 1992). Peripheral pathways, normally initiated by oxygenases, convert xenobiotics (e.g., pesticides and PAHs) into various metabolic intermediates, such as catechol and protocatechuate (Peng et al., 2008; Seo et al., 2009) (Figure 21.3). In central pathways, the intermediates undergo further metabolism to enter the TCA cycle. Application of proteomics provides insight into global cellular responses to xenobiotics and networks among diverse metabolic pathways, which is difficult to accomplish via stepwise descriptions of the individual reactions (see Figure 21.3 and Table 21.6). Catabolic enzymes are responsive to carbon sources. The -ketoadipate pathway, for example, is a chromosomally

Table 21.6  Applications of Proteomics in Bacterial Biodegradation of Pesticides and Organic Pollutants Bacteria

Substrates

Proteins

References

Benzoate, phydroxybenzoate

BenA, CatA PcaG, PcaB

Park et al. (2006)

Tetracycline (N)

OmpA38, OmpA32, CarO, OmpW

Yun et al. (2008)

Benzoate Aniline, succinate Aniline

New catechol 1,2-dioxygenase CatA(3) 48 aniline-induced proteins Malate dehydrogenase Putative ABC transporter Putative hydrolase Catechol 1,2-dioxygenases (CDI1 and CDI2) pcaG and pcaH

Yoon et al. (2007) Kim et al. (2004a) Kim et al. (2002)

OmpA-like protein Na()/H() antiporter ABC type sugar transport system Benzoate dioxygenase Phenol hydroxylase

Pessione et al. (2003)

Pseudomonadales Acinetobacter baumannii DU202

Acinetobacter lwoffii K24

p-Hydroxybenzoate Acinetobacter radioresistens S13

Acetate, benzoate, phenol

Single or mixtural substrates

Kahng et al. (2002)

Mazzoli et al. (2007)

Acinetobacter radioresistens

Benzoate, phenol

Phenol hydroxylase (PH) Benzoatedioxygenase(BD) cis-1,2-Dihydroxycyclohexa-3,5-diene-1carboxylatedehydrogenase (D)

Giuffrida et al. (2001)

Acinetobacter sp. KS-1

Benzoate

Catechol 1,2-dioxygenase

Kim et al. (2003)

Pseudomonas sp. DU102

Benzoate, phydroxybenzoate, vanillin

Protocatechuate 3,4-dioxygenase Catechol 1,2-dioxygenase Toluate 1,2-dioxygenase p-Hydroxybenzoate hydroxylase

Kim et al. (2007c)

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Table 21.6  (Continued) Bacteria

Substrates

Proteins

References

Pseudomonas alcaligenes NCIMB 9867 (strain P25X)

Gentisate

Sigma 54 rpoN

Zhao et al. (2005)

Pseudomonas alcaligenes NCIB 9867

Gentisate

Stress proteins

Zhao et al. (2004)

Pseudomonas sp. K82

Aniline, 3-methylaniline, 4-methylaniline, benzoate, p-hydroxybenzoate

Catechol 2,3-dioxygenase (CD2,3) Catechol 1,2-dioxygenase Protocatechuate 4,5-dioxygenase

Kim et al. (2004c)

Pseudomonas sp. strain phDV1

Glucose, phenol

Tsirogianni et al. (2004)

Phenol

19 proteins depending on the growth substrate 10 enzymes involved in the phenol degradation 19 inner membrane 10 outer membrane Phenol-inducible membrane transporters 49 proteins

Tsirogianni et al. (2006)

Pseudomonas putida KT2442

2-Chlorophenol

Chemical stress proteins

Lupi et al. (1995)

Pseudomonas putida F1

Toluene, phenol

10 Group T proteins 17 Group P proteins 1 Group M protein

Reardon and Kim (2002)

Pseudomonas putida KT2440

Chlorophenoxy herbicides Phenol, succinate, pyruvate

3,5-Dichlorocatechol DCC Transport, detoxification, stress response Amino acid, energy, carbohydrate, nucleotide metabolism SodM, SodF, AhpC Amino acids ABC transporters Ribose ABC transporter Sulfate ABC transporter Stress proteins Energy metabolic enzymes Alkyl hydroperoxide reductase (AphC) Benzoate dioxygenase (BenA, BenD) Catechol 1,2-dioxygenase (CatA) Protocatechuate 3,4-dixoygenase (PcaGH) -Ketoadipyl CoA thiolase (PcaF) 3-Oxoadipate enol-lactone hydrolase (PcaD) -Ketoacyl CoA thiolase (PhaD) Ring-opening enzyme (PhaL), 4-Hydroxyphenyl-pyruvate dioxygenase (Hpd) Homogentisate 1,2-dioxygenase (HmgA) Oxidative stress response (AhpC, SodB,Tpx, Dsb) General stress response (UspA, HtpG, GrpE, Tig) Energetic metabolism (AcnB, AtpH, Fpr, AceA, NuoE, MmsA-1) Fatty acid biosynthesis (FabB, AccC-1, FabBx1) Inhibition of cell division (MinD) Cell envelope biosynthesis (LpxC, VacJ, MurA) Transcription regulation (OmpR, Fur) Transport of small molecules (TolC, BraC, AotJ, AapJ, FbpA, OprQ) Downregulated proteins (PurM, PurL, PyrH, Dcd, FliC)

Benndorf et al. (2006) Kurbatov et al. (2006)

Glucose, phenol

Methyl tert-butyl ether Tetracycline, phenol

Benzoate, phydroxybenzoate, vanilline, phenylethylamine

Phenol

Benzoate, succinate

Ferric uptake regulator (Fur) Superoxide dismutases (Sod)

Papasotiriou et al. (2008)

Krayl et al. (2003) Yun et al. (2006)

Kim et al. (2006)

Santos et al. (2004)

Heim et al. (2003) (Continued)

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Table 21.6  (Continued) Bacteria

Substrates

Proteins

References

Pseudomonas putida P8

Benzoate, succinate

Eight catabolic pathway enzymes Detoxification protein Stress response proteins Carbohydrate Amino acid/protein and energy metabolism Cell envelope and cell division

Cao et al. (2008)

Pseudomonas sp. M1

Phenol, pyruvate, succinate

Six enzymes of the phenol catabolic pathway Phenol-induced stress proteins Transport proteins

Santos et al. (2007)

Pseudomonas putida DOT-TIE

Toluene

ttgDEF, ttgGHI

Segura et al. (2005)

Benzoate, penol, 4-cresol, gentisate, resorcinol

-Ketoadipate pathway enzymes Mycothiol-dependent gentisate pathway enzymes hydroxyquinol 1,2-dioxygenase Maleylacetate reductase Novel proteins (NCgl0524, NCgl0525, NCgl0527)

Qi et al. (2007)

TodX, TodR

Wang et al. (2000)

Actinomycetales Corynebacterium glutamicum ATCC 13032

Mycobacterium sp. strain PYR-1

Toluene

Mycobacterium sp. JS14.

Fluoranthene

PAH ring-hydroxylating dioxygenase 2,3-Dihydroxybiphenyl 1,2-dioxygenase trans-2-Carboxybenzalpyruvate hydratase Catalase Superoxide dismutase Chorismate synthase Nicotine-nucleotide phosphorylase

Lee et al. (2007)

Mycobacterium vanbaabenii PYR-1

Fluoranthene Pyrene

53 enzymes Ring-hydroxylating oxygenase (NidAB2, PhtAaAb) dihydrodiol dehydrogenase Ring cleavage dioxygenase Catalase-peroxidase Putative monooxygenase Dioxygenase small subunit Naphthalene-inducible dioxygenase small subunit Aldehyde dehydrogenase

Kweon et al. (2007) Kim et al. (2007a)

Pyrene, pyrene-4,5-quinone (PQ), phenanthrene, anthracene, fluoranthene

Kim et al. (2004b)

Mycobacterium vanbaabenii 6PY1

Pyrene, phenantherene

Pyrene-induced proteins

Krivobok et al. (2003)

Mycobacterium sp. strain KMS

Pyrene, pyrene-4,5-dione

Aromatic ring-hydroxylating dioxygenase 4Fe-4S ferredoxin Rieske (2Fe-2S) region Dihydrodiol dehydrogenase Oxidoreductase

Liang et al. (2006)

Acetonitrile

Acetonitrile hydratase (ANHase)

Okamoto and Eltis (2007)

Actinobacteria Rhodococcus sp. RHA1

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Table 21.6  (Continued) Bacteria

Substrates

Proteins

References

Rhodococcus jostii RHA1

Benzene, biphenyl, ethylbenzene, styrene

22 proteins (Bph or Etb pathway enzymes)

Patrauchan et al. (2008)

Rhodococcus sp. TFB

Phthalate, tetralin, naphthalene

14 proteins (pathway enzymes) Chaperonins

Tomás-Gallardo et al. (2006)

Rhodococcus sp. strain RHA1

Phenylacetic acid (PAA)

29 proteins identified

Navarro-Llorens et al. (2005)

Burkholderiales Comamonas CNB-1

4-Chloronitrobenzene

Zhang et al. (2009)

Burkholderia xenovorans LB400

4-Chlorobiphenyl, biphenyl 4-Chlorobenzoate Succinate, benzoate, biphenyl

DnaK, GroEL, AhpC, Bph BenD, CatA, DnaK, HtpG Benzoyl-coenzyme A (CoA) pathway enzymes

Agulló et al. (2007) Martínez et al. (2007) Denef et al. (2005)

2,4Dichlorophenoxypropionic acid, 2,4-dichlorophenol, 3,5-dichlorocatechol

Chlorocatechol 1,2-dioxygenases DnaK, AhpC

Benndorf and Babel (2002)

2,4Dichlorophenoxypropionic acid, 2,4dichlorophenoxyacetic acid

Chlorocatechol 1,2-dioxygenase TfdC(II) Chloromuconate cycloisomerase TfdD Tu (TufA), AhpC, SodA

Benndorf et al. (2004)

Succinate, benzoate

311 proteins exhibiting marked differences

VerBerkmoes et al. (2006)

cis-1,2-Dichloroethylene (cis-DCE)

EchA, EF-Ts, 50S ribosomal subunits L7/L12/L32/L29 Cysteine synthase A, glycerophosphodiester phosphodiesterase, iron superoxide dismutase, etc.

Lee et al. (2006)

Paracoccus denitrificans

Azide

Eight proteins

Bouchal et al. (2004)

Dehalococcoides sp. strain CBDB1

Halogenated compounds, 1,2,4-trichlorobenzene

Chlorobenzene reductive dehalogenase Formate dehydrogenase Transporter subunits Putative S-layer protein

Adrian et al. (2007)

Rhodopseudomonas palustris

p-Coumarate, benzoate, succinate

More than 1600 proteins

Pan et al. (2008)

Anaerobic benzene communities

Benzene, benzoate

Enoyl-CoA hydratase

Benndorf et al. (2009)

Alcaligenes Delftia acidovorans MC

Proteobacteria Rhodopseudomonas palustris Chlamydiales Escherichia coli

Anaerobic bacteria

(Continued)

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Bacteria

Substrates

Proteins

References

Strain EbN1

Toluene, ethylbenzene

Ebd proteins, Apc proteins

Kühner et al. (2005)

Aromatoleum aromaticum strain EbN1

22 different substrates

199 degradation pathway enzymes

Wöhlbrand et al. (2007)

p-Ethylphenol

PchCF; ChnA, EbA309, XccABC, TioL

Wöhlbrand et al. (2008)

Figure 21.3  Schematic of cellular functional systems involved in fluoranthene metabolism. Differential expression of proteins is related to catabolism pathway, TCA cycle, pentose phosphate pathway, fatty acid metabolism, nucleotide and amino acids biosynthesis, polysaccharide biosynthesis, etc. Some essential proteins (e.g., ABC-transporter, porins, regulator, chaperones) are differentially expressed in response to fluoranthene in JS14.

encoded, convergent pathway by which aromatic compounds are converted into protocatechuate and catechol. This pathway is widely utilized in soil bacteria and fungi (Harwood and Parales, 1996). The four major substrate-dependent catabolic pathways identified in Corynebacterium glutamicum ATCC 13032 are the catechol and protocatechuate branches of the -ketoadipate pathway, the mycothiol-dependent gentisate pathway, and the hydroxyquinol pathway in the cells grown on benzoate, phenol, 4-cresol, gentisate, and resorcinol (Qi et al., 2007). Hydroxyquinol 1,2-dioxygenase (Hyd/NCgl1113) and maleylacetate reductase (TdfF/NCgl1112) are involved in the hydroxyquinol pathway. Kim et al. (2003) compared the metabolic pathways in benzoate- and succinate-cultured Acinetobacter sp. KS-1. Of the 18 proteins induced with benzoate, two benzoate-degrading enzymes (catechol 1, 2-dioxygenase and -ketoadipate succinyl-CoA transferase) were identified, suggesting that benzoate degrades via

the -ketoadipate pathway in the strain KS-1. Analysis of N-terminal and internal amino acid sequences showed that this catechol 1,2-dioxygenase is highly homologous to the catechol 1,2-dioxygenase of Acinetobacter radioresistens. Two protocatechuate 3,4-dioxygenase subunits, PcaG and PcaH, have been identified in the catabolism of p-hydroxybenzoate in Acinetobacter lwoffii K24 (Kahng et al., 2002). The sequence analyses of the two subunits revealed their high similarity with PcaH and PcaG of Pseudomonas marginata or Pseudomonas cepacia in which the protocatechuate and catechol branches are established, signifying the coexistence of the two branches of the -ketoadipate pathway (Kahng et al., 2002). The benzoate-induced proteome in A. lwoffii K24 shows evidence of multiple catechol branches in the -ketoadipate pathway; a new catechol 1,2-dioxygenase CatA(3) is induced by benzoate, whereas CatA(1) and CatA(2) are expressed under the aniline culture condition (Yoon et al., 2007).

Chapter | 21  Proteomics in Pesticide Toxicology

Eighty unique proteins are differentially expressed in P. putida KT2440 in response to six different organic compounds (Kim et al., 2006). The metabolic pathways of benzoate, p-hydroxybenzoate, and vanilline catalyzed by substrate-specific dioxygenases (e.g., benzoate dioxygenases, catechol 1,2-dioxygenase, and protocatechuate 3,4-dixoygenase) converge into the -ketoadipate degradation pathway, in which -ketoadipyl CoA thiolase and 3-oxoadipate enollactone hydrolase participate (Kim et al., 2006). Pseudomonas sp. DU102 cultured on benzoate or p-hydroxybenzoate shows not only the  and  subunits of protocatechuate 3,4-dioxygenase but also catechol 1,2-dioxygenase, both of which are responsible for ortho cleavage of the aromatic compounds. Toluate 1,2-dioxygenase and p-hydroxybenzoate hydroxylase are also found in Pseudomonas sp. DU102 (Kim et al., 2007c). The proteome of Pseudomonas sp. K82 suggests three main metabolic pathways in which the catechol 2,3-dioxygenase and catechol 1,2-dioxygenase (ketoadipate) play major and secondary roles, respectively, in the metabolism of aniline and its analogues and in which the protocatechuate 4,5-­dioxygenase is mainly involved in p-hydroxybenzoate metabolism (Kim et al., 2004c). Mycobacterium vanbaalenii PYR-1 is the first bacterium isolated by virtue of its ability to metabolize pyrene (Moody et al., 2004). The five tandem genes encode phthalate dioxygenase large () subunit (phtAa), small subunit () (phtAb), phthalate dihydrodiol dehydrogenase (phtB), phthalate dioxygenase ferredoxin subunit (phtAc), and phthalate dioxygenase ferredoxin reductase (phtAd) in M. vanbaalenii PYR-1 (Moody et al., 2004; Stingley et al., 2004). Kim et al. (2004b) detected more than 1000 proteins from M. vanbaalenii PYR-1 incubated with pyrene, pyrene-4,5-quinone, phenanthrene, anthracene, or fluoranthene. Among the identified PAH-induced proteins are catalase-peroxidase, a putative monooxygenase, dioxygenase  subunit, naphthalene-inducible dioxygenase  subunit, and an aldehyde dehydrogenase. Kim et al. (2007b) constructed a pyrene degradation network through o-phthalate and the -ketoadipate pathways in M. vanbaalenii PYR-1. Of the 18 upregulated enzymes, the terminal subunits of ring-hydroxylating oxygenase, dihydrodiol dehydrogenase, and ring cleavage dioxygenase are detected only in pyrene-grown cells. Kweon et al. (2007) integrated the metabolic information with the genomic and proteomic results from M. vanbaalenii PYR-1 and proposed the degradation pathways of fluoranthene, which consists of 18 enzymatic steps via 9-fluorenone-1-carboxylic acid and phthalate with the initial ring-hydroxylating oxygenase (NidA3B3) oxidizing fluoranthene to fluoranthene cis-2, 3-dihydrodiol. Lee et al. (2007) examined proteins responsible for fluoranthene catabolism in Mycobacterium rufum JS14 efficiently degrading fluoranthene. The approximately 25 upregulated proteins associated with fluoranthene metabolism include several Bph-degrading enzymes, such as 2,3-dihydroxybiphenyl 1,2-dioxygenase (BphC), 4-hydroxy-2-oxovalerate aldolase (HOVA aldolase and

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BphF), biphenyl dioxygenase (BphG), response regulator of biphenyl metabolism gene cluster (BphT), and 1-hydroxy2-naphthoate dioxygenase. Two key proteins, ARH dioxygenase  and  subunits, were found to be overexpressed in response to fluoranthene in M. rufum JS14. Proteomic analyses have also revealed enzymes necessary for biphenyl (Bph) and ethylbenzene (Etb) catabolism (i.e., Bph and Etb pathways) in Rhodococcus jostii RHA1 (Patrauchan et al., 2008). Of the 151 identified proteins, 22 Bph/Etb proteins are among the most abundant in biphenyl-, ethylbenzene-, benzene-, and styrene-grown cells. Both Bph and Etb enzymes and at least two sets of lower Bph pathway enzymes were found in the cells grown on biphenyl, ethylbenzene, or benzene. However, no Etb enzymes and only one set of lower Bph pathway enzymes are expressed in the styrene-grown cells. In other words, both Bph and Etb dioxygenases preferentially catalyze transformation of biphenyl, whereas only Etb dioxygenase catalyzes transformation of styrene. Gene disruption confirmed that styrene and benzene are degraded via meta and ortho cleavage, respectively (Patrauchan et al., 2008). A similar study identified 14 enzymes from phthalategrown Rhodococcus sp. TFB allowing a complete delineation of the catabolic pathway of phthalate to the TCA cycle via intradiol (ortho) cleavage of protocatechuate (TomásGallardo et al., 2006). Burkholderia xenovorans LB400 (formerly known as B. fungorum) (Goris et al., 2004), an aerobic PCB degrader, degrades a wide range of PCBs (Bedard et al., 1986). Genomic and proteomic studies of the succinate-, benzoate-, or biphenyl-grown cells show three benzoate pathways (a catechol ortho cleavage and two benzoyl-CoA pathways) and the C1 metabolic pathway depending on the growth substrate and phase (Denef et al., 2004, 2005). Exposures to 4-chlorobiphenyl inhibited the growth of LB400 on glucose; the cells exhibit irregular outer membranes, a larger periplasmic space, and electron-dense granules in the cytoplasm. Chlorobenzoates induce the enzymes BenD and CatA in benzoate and catechol catabolic pathways in B. xenovorans LB400 (Martínez et al., 2007). Harwood et al. (1998) and Heider and Fuchs (1997) outlined the anaerobic pathways that allow bacteria to utilize aromatics even in the absence of oxygen. Anaerobic aromatic metabolism undertakes a set of oxygen-free reactions and forms different central intermediates (e.g., benzoyl-CoA) for aromaticity cleavage; notably, the aromatic ring is reduced rather than oxidized. Key intermediates in anaerobic aromatic metabolism include benzoyl-CoA and compounds having at least two meta-positioned hydroxyl groups (e.g., resorcinol, phloroglucinol, and hydroxyhydroquinone) (Boll, 2005). Many anaerobic respiration pathways of aromatics and some of the key enzymes have been characterized in detail (Adrian et al., 2000; Bunge et al., 2003; Foght, 2008). Aromatic growth substrates such as toluene, phenol, cresols, xylenes, ethylbenzene, and benzoate analogues are channeled to the central intermediate benzoyl-CoA

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prior to dearomatization and ring cleavage (Boll, 2005; Klein, 2000; Schink and Friedrich, 2000). Rabus (2005) summarized current advances in multiple respiratory complexes and an extensive regulatory network in the anaerobic bacterium Azoarcus sp. EbN1. Functional genomic analysis revealed 10 anaerobic and 4 aerobic aromatic degradation pathways in the strain EbN1. The strain EbN1 degrades toluene and ethylbenzene anaerobically via completely different pathways. The global expression patterns of anaerobically grown cells show specific induction of two toluene-related operons (bss and bbs) in toluene-adapted cells, whereas Ebd proteins (encoding subunits of ethylbenzene dehydrogenase) are formed in ethylbenzene- but not in acetophenone-adapted cells (Kühner et al., 2005). Wöhlbrand et al. (2008) reported utilization of p-ethylphenol by an anaerobic ethylbenzene pathway under anoxic conditions in Aromatoleum aromaticum EbN1, where the induced proteins include a p-cresol methylhydroxylase-like protein (PchCF), two predicted alcohol dehydrogenases (ChnA and EbA309), a biotin-dependent carboxylase (XccABC), and a thiolase (TioL). An integrated transcriptomics and quantitative proteomics study characterized anaerobic catabolism of p-coumarate in Rhodopseudomonas palustris in which p-coumarate is converted to benzoyl-CoA and then degraded further via a known aromatic ring reduction pathway by a non--­oxidation route (Pan et al., 2008).

21.4.2  Network of Catabolism and Central Metabolism Xenobiotics-degrading bacteria can utilize pesticides for carbon and energy sources (see Table 21.6). The network of catabolism and central metabolism is directly related to efficiency of substrate utilization, particularly in nutrientdeficient environments. Metabolomes and proteomes dyna­ mically respond to substrates and chemical stressors (see Chapter 22). Proteome analyses of fluoranthene-grown M. rufum JS14 cells suggested overexpression of 3-oxoadipate enol-lactone hydrolase and other TCA cycle enzymes playing a role to sustain the TCA cycle and overexpression of acetyl-CoA synthase to support the synthesis of steroids or lipid metabolism (see Figure 21.3) (Lee et al., 2007). Induction of the tyrosine biosynthesis enzymes, such as chorismate synthase and 4-hydroxyl-phenylpyruvate (4-HPP) dioxygenase, suggests that biosynthesis of aromatic amino acids (e.g., tyrosine) is altered when the strain JS14 utilizes fluoranthene. Comparative proteome profiles from C. glutamicum cultured on different aromatic compounds or glucose indicate that the central carbon metabolism varies with the substrates (Qi et al., 2007). Sixteen enzymes associated with the central metabolism are differentially expressed. Among the 16 differentially expressed proteins are isocitrate lyase in the glyoxylate shunt, citrate synthase and aconitase A in the TCA cycle, and pyruvate kinase and

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pyruvate:quinone oxidoreductase in glycolysis. Differential expression of these enzymes signifies fine-tuning of carbon flux at the phosphoenolpyruvate–pyruvate–oxaloacetate node. Proteomics analyses, complemented with genetic deletion, of C. glutamicum ATCC 13032 defined a clear link between aromatic degradation and central carbon metabolism and cell growth via the gluconeogenesis. Fructose-1,6-bisphosphatase (Fbp), a rate-limiting enzyme in gluconeogenesis, was confirmed to play a key role in aromatic assimilation by C. glutamicum (Qi et al., 2007). Increased abundance of Fbp was previously observed in a 54-deficient mutant of P. alcaligenes NCIMB 9867 under gentisate induction (Zhao et al., 2005). There is no homolog of σ54 in the C. glutamicum genome (Pátek, 2005). Proteome changes imply systemic responses in the metabolically engineered Escherichia coli, including enhanced aerobic degradation of cis-1,2-dichloroethylene (cis-DCE), enhanced synthesis of glutathione to reduce toxicity from cis-DCE epoxide, and repression of fatty acid synthesis, gluconeogenesis, and the TCA cycle (Lee et al., 2006a). Among the identified proteins from M. vanbaalenii PYR-1 cells grown on pyrene, pyrene-4,5-quinone, phenanthrene, anthracene, and fluoranthene are those involved in carbohydrate metabolism (e.g., enolase, 6-phosphogluconate dehydrogenase, indole-3-glycerol phosphate synthase, and fumarase), DNA translation, Hsps, and energy production (e.g., ATP synthase) (Kim et al., 2004b). A proteomics study identified eight catabolic enzymes involved in both the ortho cleavage (CatB, PcaI, and PcaF) and the meta cleavage (DmpC, DmpD, DmpE, DmpF, and DmpG) pathways for benzoate catabolism in P. putida P8 exposed to high benzoate concentrations. In addition, some of the other 28 differentially expressed proteins are involved in (1) detoxification and stress response (e.g., AhpC, ATPase-like ATPbinding region, putative DNA-binding stress protein, SodB, and catalase/peroxidase HPI); (2) carbohydrate, amino acid/ protein, and energy metabolism (e.g., isocitrate dehydrogenase, SucC, SucD, AcnB, GabD, ArcA, ArgI, Efp, and periplasmic binding proteins of ABC transporters); and (3) cell envelope and cell division (e.g., bacterial surface antigen family protein and MinD) (Cao and Loh, 2008).

21.4.3  Bacterial Cell Membrane Proteins Degradation of pesticides in bacteria is a complex physiological phenomenon requiring the first, critical step of transport of the chemicals through the membrane into the cells. Cell membranes contain lipids, enzymes, structural proteins, recognition proteins, receptors, and transporters. Detoxification is another requirement of bacterial degradation of pesticides. Studies of the bacterial cell membrane proteome have helped to elucidate the role of the membrane in communication between internal and external environments and stress responses. Cell membrane proteins are of great significance for the biodegradation of pollutants; many factors affect cell membrane proteins (Pessione

Chapter | 21  Proteomics in Pesticide Toxicology

et al., 2003; Richins et al., 1997; Wang et al., 1995). Upon exposure, the cell membrane alters its lipid composition, hydrophobicity, and levels of specific proteins. Study of the bacterial membrane proteome is of growing interest in the research of nutrient transport and processing. Unfortunately, there are few publications on applications of proteomics to study bacterial cell membrane proteins in the biodegradation of pesticides and organic pollutants. Upon exposure to aromatics, the membrane proteome in A. radioresistens S13 includes an Na/H antiporter and ABC-type sugar transporters; the former is likely involved in the regulation of intracellular pH and the latter is probably involved in capsular polysaccharide translocation (Pessione et al., 2003). Other overexpressed cell membrane proteins in aromatic-grown cells include: (1) an OmpA-like protein on the outer membrane that enhances bioavailability of hydrocarbons; (2) another outer membrane protein, the trimeric porin of the PhoE family, which facilitates the transport of anions, particularly phosphate; and (3) two glycosyl transferases that are probably associated with capsules and/or lipopolysaccharide biosynthesis (Pessione et al., 2003). Comparison of the membrane subproteomes during growth of Pseudomonas sp. phDV1 on glucose or phenol revealed 19 inner membrane proteins and 10 outer membrane proteins. Two membrane proteins are only expressed in the presence of phenol, and one of them may function as an aromatic compound-specific porin (Papasotiriou et al., 2008). Ralstonia eutropha seems to overcome the formic acid toxicity by increasing ion transporters and formic acid metabolism catalyzed by formate hydrogenylase, a membrane enzyme (Lee et al., 2006b).

21.4.4  Bacterial Stress Responses and Adaption Exposure to toxic chemicals triggers a cascade of cellular responses that allow the bacterium to defend, detoxify, and adapt to the particular environment or stressor (Hightower, 2003; Ram et al., 2005; Storz and Hengge-Aronis, 2000). Chemical-induced stresses include heat/cold shock, oxidative stress, and general stress responses. Proteomic analyses are arguably the best way to provide a comprehensive overview of the adaptation mechanisms for bacterial response to pesticides. A large number of phenol-induced stress proteins in P. putida KT2440 are classified as antioxidant enzymes, heat shock proteins, and chaperones (Santos et al., 2004). Four stress-related proteins—the Hsp GroES, cold shock protein CspA2, translational elongation factor EF-Tu-1, and the xenobiotic reductase (XenA)—are up-expressed in toluene-grown P. putida DOT-T1E cells (Segura et al., 2005). The overproduction of these stress-related proteins indicates that toluene causes stresses, which is in agreement with its toxic character. Global proteome analyses showed up-expression of alkylhydroperoxide reductase

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C (AhpC) and superoxide dismutase (SodM and SodF) in P. putida KT2440 exposed to methyl tert-butyl ether (Krayl et al., 2003). Alkyl hydroperoxide reductase (AphC) is induced by all aromatic compounds tested (e.g., benzoate, p-hydroxybenzoate, vanilline, phenylethylamine, and phenylalanine) in P. putida KT2440 (Kim et al., 2006). The biosynthesis of ferric uptake regulator-dependent proteins is repressed by chlorophenoxy herbicides and their initial metabolites in P. putida KT2440 due to uncoupling of oxidative phosphorylation (Benndorf et al., 2006). Functional proteome analyses revealed induction of many heat shock proteins in response to various stimuli (Chen et al., 2000; Giard et al., 2002; Monahan et al., 2001). Examples of bacteria studied on production of heat shock proteins to alleviate harsh stimuli are Pseudomonas (Kim et al., 2006; Zhao et al., 2007), Methylocystis (Uchiyama et al., 1999), and Burkholderia species (Cho et al., 2000). When P. alcaligenes P25X cells are cultured at 32 or 42°C in the presence or absence of gentisate as a stressor, 19 heat shock proteins are differentially expressed. These heat shock proteins are categorized into six classes: Hsp45, Hsp60, Hsp70, Hsp90, Hsp100, and sHsp. Overproduction of enzymes in the biosynthesis of glutathione, formate detoxification enzymes, and heat shock proteins is a defense mechanism to alleviate toxic effects of formic, acetic, propionic, and levulinic acids in R. eutropha (Lee et al., 2006b, 2009). Catalase and superoxide dismutase are up-expressed to relieve oxidative stresses in M. rufum JS14 (Lee et al., 2007). 4-Chlorobenzoate (4-CBA) induces the expression of the enzymes BenD and CatA of benzoate and catechol catabolic pathways in B. xenovorans LB400. The induction of molecular chaperones DnaK and HtpG by 4-CBA indicates that the exposure of 4-CBA constitutes a stressful condition for the strain LB400 (Martínez et al., 2007). Several eleva­ ted proteins in B. xenovorans LB400 cells include general stress proteins in response to benzoate and biphenyl (Denef et al., 2005), chaperones DnaK and GroEL in response to 4-chlorobiphenyl or biphenyl, and alkyl hydroperoxide reductase AhpC in response to biphenyl (Agulló et al., 2007). Profiles of stress proteins are generally chemical specific.

Conclusion Proteomics is an emerging discipline in pesticide toxicology. It has gained acceptance in numerous areas of pesticide research, such as pesticide metabolism and mechanisms of toxicity. Proteomics has already begun to enhance understanding of mechanisms of pesticide resistance, mechanisms of toxic action, mode of action, and biodegradation of pesticides, and it has aided in the discovery of new effective and safe pesticides and in identification of biomarker proteins. However, there are challenges in proteomics methodologies, including extending the dynamic range to cover low- and high-abundant proteins and performing efficient protein quantitation and data mining. The

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diversity of potential and combinatorial post-translational modifications adds additional complexity in proteomics studies. Integration of proteomics with bioassays and other omics, such as genomics, transcriptomics, and metabolomics, can certainly provide specific, comprehensive, and in-depth knowledge in pesticide toxicology. A major challenge has arisen to integrate proteomics with other omics technologies, particularly metabolomics, in which low-molecular-weight primary and secondary metabolites are key players in biodegradation. Despite the reported excellent success of proteomics in common model organisms, analyses of the complex proteomes and characterization of functional proteomes in species beyond the model species require much effort. The development of protein and genomic databases will facilitate the application of proteomics for other species. Streamlining protein preparation and fractionation with a suitable analytical technique (e.g., MS and NMR) is essential for extending the potential of proteomics to pesticide toxicology. Although MS-based approaches are very powerful for qualitative metaproteome investigations, there is a great need to develop and demonstrate improved approaches for quantitative measurements. Furthermore, the ability to characterize protein post-translational modifications is essential for a more comprehensive understanding of how a species of interest regulates proteins for functionality and toxic responses to pesticides. For microbial remediation purposes, proteo-arrays can detect binding of specific inhibitors or ligands with dioxygenases or monooxygenases. Key catabolic enzymes can be profiled to elucidate the network with neighboring proteins based on qualitative and quantitative estimation during in situ bioremediation. However, to date, no study has identified the global interactions involving proteins (i.e., interactomics) in an organism during bioremediation processes. In addition, the impact of single nucleotide polymorphisms on proteome analyses by MS requires further exploration. Further improvements in MS technology and methodology are of significance in life sciences, including pesticide toxicology. Finally, a need in MS-based proteomics is to make use of the enormous amount of data being generated. To analyze proteome data, one must understand the analytical procedures used to obtain the data and the statistical principles underlying multiple dimensional data. Proteomics is becoming an indispensable tool in pesticide toxicology.

Acknowledgments This work was supported in part by grants from the State of Hawaii, the Department of Agriculture, the U.S. Fish and Wildlife Service, USDA Tropical and Subtropical Agricultural Research awards, U.S. EPA award 98951201-1, U.S. NRL award N00173-05-2-C003, and Hawaii Energy and Environmental Technologies Initiative Award N00014-09-1-0709. We thank Margaret R. Ruzicka for a comprehensive review of this chapter.

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Fe-only hydrogenase in Clostridium acetobutylicum. Appl. Environ. Microbiol. 69(3), 1542–1547. Westergren-Thorsson, G., Malmström, J., and Marko-Varga, G. (2001). Proteomics—The protein expression technology to study connective tissue biology. J. Pharm. Biomed. Anal. 24(5-6), 815–824. Wilkins, M. R., Williams, K. L., Appel, R. D., and Hochstrasser, D. (1997). “Proteome Research, New Frontiers in Functional Genomics,” Springer-Verlag, Berlin. Wöhlbrand, L., Kallerhoff, B., Lange, D., Hufnagel, P., Thiermann, J., Reinhardt, R., and Rabus, R. (2007). Functional proteomic view of metabolic regulation in “Aromatoleum aromaticum” strain EbN1. Proteomics 7(13), 2222–2239. Wöhlbrand, L., Wilkes, H., Halder, T., and Rabus, R. (2008). Anaerobic degradation of p-ethylphenol by “Aromatoleum aromaticum” strain EbN1: pathway, regulation, and involved proteins. J. Bacteriol. 190(16), 5699–5709. Yang, Y., Zhang, S., Howe, K., Wilson, D. B., Moser, F., Irwin, D., and Thannhauser, T. W. (2007). A comparison of nLC-ESI-MS-MS and nLC-MALDI-MS-MS for GeLC-based protein identification and iTRAQ-based shotgun quantitative proteomics. J. Biomol. Tech. 18(4), 226–237. Yates, J., Ruse, C. I., and Nakorchevsky, A. (2009). Proteomics by mass spectrometry: Approaches, advances, and applications. Annu. Rev. Biomed. Eng. 11, 49–79. Yoon, Y. H., Yun, S. H., Park, S. H., Seol, S. Y., Leem, S. H., and Kim, S. I. (2007). Characterization of a new catechol branch of the β-ketoadipate pathway induced for benzoate degradation in Acinetobacter lwoffii K24. Biochem. Biophys. Res. Commun. 360, 513–519. Yun, S. H., Kim, Y. H., Joo, E. J., Choi, J. S., Sohn, J. H., and Kim, S. I. (2006). Proteome analysis of cellular response of Pseudomonas putida KT2440 to tetracycline stress. Curr. Microbiol. 53(2), 95–101. Yun, S. H., Choi, C. W., Park, S. H., Lee, J. C., Leem, S. H., Choi, J. S., Kim, S., and Kim, S. I. (2008). Proteomic analysis of outer membrane proteins from Acinetobacter baumannii DU202 in tetracycline stress condition. J. Microbiol. 46(6), 720–727. Zhang, Q., and Riechers, D. E. (2004). Proteomic characterization of herbicide safener-induced proteins in the coleoptile of Triticum tauschii seedlings. Proteomics 4(7), 2058–2071. Zhang, Q., Xu, F., Lambert, K. N., and Riechers, D. E. (2007). Safeners coordinately induce the expression of multiple proteins and MRP transcripts involved in herbicide metabolism and detoxification in Triticum tauschii seedling tissues. Proteomics 7(8), 1261–1278. Zhang, Y., Wu, J. F., Zeyer, J., Meng, B., Liu, L., Jiang, C. Y., Liu, S. Q., and Liu, S. J. (2009). Proteomic and molecular investigation on the physiological adaptation of Comamonas sp. strain CNB-1 growing on 4-chloronitrobenzene. Biodegradation 20, 55–66. Zhao, B., Yeo, C. C., Lee, C. C., Geng, A., Chew, F. T., and Poh, C. L. (2004). Proteome analysis of gentisate-induced response in Pseudomonas alcaligenes NCIB 9867. Proteomics 4(7), 2028–2036. Zhao, B., Yeo, C. C., and Poh, C. L. (2005). Proteome investigation of the global regulatory role of sigma 54 in response to gentisate induction in Pseudomonas alcaligenes NCIMB 9867. Proteomics 5, 1868–1876. Zhao, B., Yeo, C. C., Tan, C. L., and Poh, C. L. (2007). Proteome analysis of heat shock protein expression in Pseudomonas alcaligenes NCIMB 9867 in response to gentisate exposure and elevated growth temperature. Biotechnol. Bioeng. 97(3), 506–514. Zolla, L., Rinalducci, S., Antonioli, P., and Righetti, P. G. (2008). Proteomics as a complementary tool for identifying unintended side effects occurring in transgenic maize seeds as a result of genetic modifications. J. Proteome Res. 7(5), 1850–1861.

Chapter 22

Metabolomics in Pesticide Toxicology Young Soo Keum1, Jeong-Han Kim1 and Qing X. Li2 1 2

Seoul National University, Seoul, Korea University of Hawaii, Honolulu, Hawaii

22.1  Introduction The recent era of molecular biology can be characterized by the rapid accumulation of entire genomic contexts of various organisms. Currently, genomes of more than 400 species are fully sequenced, including humans, several mammals, fishes, plants, and microorganisms. The flourishing genetic information and advances in analytical tools opened the new disciplines of “systems biology” and “omics” (Figure 22.1). Such research interests are evidenced by a dramatic increase in the number of publications and patents in metabolomics and proteomics since 1990 (Figure 22.2). Elementary reactions or unit biochemical pathways are simply connected in conventional research; however, biological systems (biosystems) are highly interwoven networks of unit reactions at levels of nucleotides, proteins, and biomolecules. Such complexity is further complicated by hierarchical regulations and molecular signaling. These complexities necessitate sophisticated evaluation of whole interaction in biosystems to which “omics” approaches contribute well. This chapter discusses the definition, historical evolution, methodologies, and applications of metabolomics in DNA

Genomics

RNAs

Transcriptomics

Proteins

Proteomics

Metabolites

Metabolomics

Figure 22.1  Schematic diagram of multi-omics and their target biomolecules. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

pesticide toxicology. Chemical toxicology, which includes pesticide toxicology, examines dose responses to toxic chemicals (also referred to as chemical stimuli). The metabolic responses that are either direct effects (e.g., inhibition of enzymes) or a cascade of indirect regulatory circuits (e.g., receptor and agonist or antagonist events) are compared to those of controls. These interactions result in differential expression of genes, proteins, and metabolites, which are also accompanied by characteristic physiological phenomena (e.g., tumor formation, metabolic anomalies, and even death). Future perspectives of metabolomics research and applications are briefly discussed. Many excellent reviews of metabolomics are available (Bundy et al., 2009; Griffin, 2003; Robertson, 2005; Simpson and McKelvie, 2009; Viant 2007, 2008, 2009).

22.2  Metabolism, metabolites, and metabolomics 22.2.1  Metabolism and Metabolites Metabolism encompasses the biological reactions and/or nonbiological reactions required for an organism to sustain life (see Chapter 38). The metabolic processes are considered either primary or secondary. Primary metabol­ ism is defined as metabolism that is indispensable to sustain the functioning of normal biological systems, whereas secondary metabolism is considered as noncritical but advantageous to enhance the tolerance and competence of organisms to external stimuli. Metabolites in primary metabolism are highly conserved throughout a wide variety of taxa, whereas secondary metabolites are produced in species-specific or organ-specific biosynthetic pathways (Luckner, 1990). Pesticide catabolism is the degradation of pesticides in organisms. Metabolism, or the fate of specific metabolites, is under the influence of various stimuli, and the test compound can trigger numerous genetic or metabolic responses. In plants, it is well-known that pesticides 627

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can modulate not only primary metabolism but also biosynthesis of secondary metabolites (Lydon and Duke, 1988). 2,3,7,8-Tetrochlorodibenzo-p-dioxin (TCDD), an impurity in the herbicide Agent Orange, induced various responses of gene elements in mice, including primary metabolism-related enzymes and regulatory elements, and stress responses (Tijet et al., 2006).

22.2.2  Metabolomics Metabolomics is the systematic study of a metabolome, the entirety of metabolites, or a set of metabolites, forming an extensive network of metabolic reactions in which one metabolite from a specific pathway will affect one or more biochemical reactions, or a comprehensive and quantitative analysis of all metabolites (Fiehn, 2001; Oliver et al., 1998). The term “metabolome” was introduced by Oliver (Oliver et al., 1998) to represent sets of all metabolites from specific biosystems. Although the term was introduced recently, analogous research has a long history; for example, metabolite profiling or fingerprinting has been applied in various biochemical and toxicological studies since the 1980s (Blanchard et al., 1985; Nicholson et al., 1985). Comparative profiling of metabolites has long been used in chemotaxonomy of microorganisms (e.g., fatty acid methyl ester profile analysis of bacterial classification). The experimental procedures used in early biochemical research fall under the definition of metabolomics, although there are distinctive differences. Current metabol­ omics addresses comprehensive sets of metabolites and

the systemwide interpretation of the corresponding data as well as specific pathways. Regarding other emerging scientific areas, there are several arguments about appropriate terminologies or terms for metabolomics and closely related research. The most representative example of the arguments is the definition of “metabolomics” versus “metabonomics,” where the distinction is mainly philosophical rather than technical (Nicholson and Lindon, 2008). Both words have been used interchangeably in many research papers (Beckonert et al., 2003; Bruder et al., 2004; Bundy et al., 2009; Cherney et al., 2007; Ekman et al., 2006; Fiehn, 2002; Griffin, 2003; Guo et al., 2009). Many publications titled “metabolomic” and “metabonomic” share similar conceptual frameworks and experimental methods. However, scientists in metabonomics argue that “metabonomics broadly aims to measure the global, dynamic metabolic response of living systems to biological stimuli or genetic manipulation while metabolomics seeks an analytical description of complex biological samples, and aims to characterize and quantify all the small molecules” (Nicholson and Lindon, 2008, p. 1054; see also Nicholson et al., 2002). In such sense, most metabolomic research can be regarded as metabon­ omics. Because such research is typically characterized by comparative studies between different stimuli (e.g., different toxicants, concentrations, and temperatures), most metabol­ omics/metabonomic studies share similar experimental protocols, including extraction of metabolites (e.g., metabolome), proper derivatization of metabolites when necessary, instrumental analyses, and structural identification of metabolites followed by data mining (Figure 22.3).

Log number of publications and patents

100000

10000

1000 Genomics Proteomics Metabolomics

100

10

2008–2009

2006–2007

2004–2005

2002–2003

2000–2001

1998–1999

1996–1997

1994–1995

1992–1993

1990–1991

Before 1990

1

Figure 22.2  Numeric overview of documents (journal articles, patents, books, and other related publications) containing the concepts of genomics, proteomics, and metabolomics. The data were compiled from an exhaustive search with SciFinder Scholar. Numbers of publication were log-scaled. The results for 2009 are as of March 2009.

Chapter | 22  Metabolomics in Pesticide Toxicology

Comprehensive mapping or profiling of metabolites is somewhat similar to research seeking to construct proteome reference maps (Lei et al., 2006; Zhan and Desiderio, 2003). In this chapter, metabolomics and metabonomics are regarded as interchangeable in that they both are defined as the comprehensive analyses of a large number of metabol­ ites, including interpretation of the functions and network of metabolites and understanding of metabolic responses and phenomena triggered by stimuli or genetic manipulations in biosystems. Therefore, metabolomics, as a major part of systems biology, is a far more profound concept and practice than simple metabolic profiling and biomarker search, although comprehensive, qualitative, and quantitative analyses of metabolites are the essential and common procedures in metabolomics studies. Unit reactions and constituents (e.g., precursors, products, and enzymes) are highly interconnected; quantitative assessment of metabolic flux should give more detailed information about the biosystems in question. Such scientific needs unravel a new part of omics (Wiechert et al., 2007), namely fluxomics as a discipline “that analyzes the fluxome as one part of systems biology and provides mathematically defined networks of metabolic reactions and their regulation” (Bornholdt, 2005, p. 449). In comparison with ordinary metabolomics, fluxomics or metabolic flux

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analysis focuses on kinetic aspects of the metabolomic changes induced by stimuli. Requisite to fluxomics is the accurate and reproducible quantitation of a metabolome. Differentiation of lipid profiles is a well-known approach in biosystems under a chemical stimulus or disease (Astarita and Piomelli, 2009; Bruder et al., 2004; Cheng et al., 2008; Davies, 2009). Indeed, “lipidomics” is an emerging field within metabolomics and is defined as a lipid-targeted metabolomics approach aimed at comprehensive analysis of lipids in biological systems (Hu et al., 2009). “Toxicometabolomics” is another branch of metabol­ omics. It involves studying toxicological responses using metabolomic approaches. Toxicometabolomics contributes valuable information for the understanding of the mechanisms of toxicity, mode of toxic action, and adverse effects of pesticides, toxicants, drugs, and bioactive natural or synthetic products.

22.3  Research methods in metabolomics According to its definition, metabolomics can theoretically quantify all metabolites in all metabolic pathways. However, it is practically impossible to cover all metabol­ ites with a single analytical method because biological

Biological sample

Identification of sample Sample pretreatment (e.g., removal of excess impurities)

Quenching, extraction (e.g., liquid nitrogen) Derivatization as needed (e.g., methoximation, silylation) Instrumental analyses (e.g., MS, NMR)

Data processing (e.g., peak alignment, statistics) Confirmatory experiments (e.g., genomics, transcriptomics)

Results Figure 22.3  A typical metabolomic workflow.

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samples usually contain a wide range of classes of metabol­ ites with extreme differences in physicochemical properties (e.g., solubility, stability, and relative abundance). For example, phosphorylated sugars are water-soluble, whereas most cell membrane constituents show a degree of hydrophobicity. To investigate all metabolites, several sets of sample processing protocols must be applied. Sample property is another important aspect to consider when evaluating a specific metabolomic workflow. Cautious and rapid handling of samples is very important in a metabol­ omics workflow because metabolic differentiation upon exposure to stimuli is rapid and can prevent representation of the metabolome as it is. Therefore, a variety of methods have been developed to meet various needs and even for the same type of samples, with each method having strengths and weaknesses. Common sample pretreatment and extraction protocols in metabolomics are discussed in the next section.

22.3.1  Sample Pretreatment and Extraction Methods The large degree of variation in test specimens and their metabolic profiles necessitates specialized sample treatment approaches. Some noteworthy approaches are highlighted in Table 22.1. In the case of environmental and field samples, precautions need to be taken and are not limited to rapid quenching and inhibition of enzymatic reactions, proper pretreatment such as tissue homogenization,

reproducible and efficient extraction of metabolites, proper cleanup of unwanted metabolites, efficient concentration of extracts, and solid identification of biological specimens. Sample pretreatment prior to extraction of metabolites is often required because an excess of specific metabol­ ites and/or contaminants can prevent proper identification of other metabolites. The cellular constituents of cultured cells are often contaminated by culture medium, leading to uninformative results. Similarly, field samples taken from soil or aquatic ecosystems can be contaminated with complex environmental constituents, leading to biased results. Careful washing with buffers or other solvents is commonly applied to remove debris and trace contaminants (Bundy et al., 2002a,b; Guo et al., 2009; Keum et al., 2008; McKelvie et al., 2009). High concentrations of urea in urine make it difficult to analyze some amino acids. Treatment of the samples with urease can ameliorate this problem (Kind et al., 2007; Kuhara, 2007). High concentrations of proteins can be removed from samples through precipitation with trichloroacetic acid. In a metabol­omics study targeting specific sets of metabolites, specialized enrichment techniques can be utilized (Carlson and Cravatt, 2007a,b). Low-molecular-weight metabolites are prone to degradation in response to environmental changes whether they are internal or external; consequently, rapid quenching of metabolic enzymes is one of the most important prerequisites for accurate analysis. Rapid freezing in liquid nitrogen, for example, is often used to completely stop metabolic events (McKelvie et al., 2009; Tuffnail et al., 2009).

Table 22.1  Examples of Extraction Methods in Metabolomic Analyses Sample

Extraction solvents

Extraction method

Reference

Bacteria

Mixture of methanol and water

Sonication

Keum et al. (2008)

Cancer cells

Multiple set of organic solvent/water

Homogenization

Sreekumar et al. (2009)

Daphnia magna

Mixture of methanol, chloroform, and water (cold)

Homogenization

Taylor et al. (2009)

Earthworms

Buffer

Sonication

McKelvie et al. (2009)

Fish (Danio rerio)

Multiple set of organic solvent/water

Homogenization

Ong et al. (2009)

Plant cells

Mixture of methanol and water

Homogenization

Sarry et al. (2006)

Plant leaves

None, in situ analyses

Cha et al. (2008)

Mixture of methanol, chloroform, and water (cold)

Homogenization

Weckwerth et al. (2004)

Rat urine

None

Centrifugation

Beckonert et al. (2003)

Sera

Methanol

Centrifugation

Zelena et al. (2009)

Culture medium, yeast

None

Centrifugation

Allen et al. (2004)

Miscellaneous animal samples

Mixture of chloroform and methanol

Homogenization

Astarita and Piomelli (2009)

Chapter | 22  Metabolomics in Pesticide Toxicology

In practice, both quenching and metabolite extraction procedures are performed simultaneously. Use of organic solvent is a representative option for quenching and metabolite extraction. Aqueous alcohols are the most common choice for polar metabolites (see Table 22.1). Several classical methods (e.g., the Folch method and the Bligh and Dyer method or its variant) are commonly used for lipid metabol­ ites. Similar extraction methods can also be used to isolate and separate polar metabolites from lipids and pigments. In comparison with liquid samples such as serum and urine, homogeneous, representative sampling is a difficult task for solids such as cells, tissues, and organs. These samples are usually finely divided or pulverized prior to extraction. Intracellular metabolites from microorganisms are extracted after cell disruption by enzymatic or physical methods, including lysozymes, bead beaters, ultrasonication, and freeze–thaw sequences. Homogenization in liquid nitrogen is a common practice for plant tissues. After extraction, concentration of extracts is usually performed. Special precautions should be taken because some metabolites such as short-chain organic acids are highly volatile and can be lost during a concentration step. Although there is limited coverage of storage methods in the literature, this is a critical component to metabol­ omics studies. Metabolomics studies often necessitate a large number of samples being stable for storage, so attention to this subject is warranted. Perhaps the most common method for storage is lyophilization after quenching with liquid nitrogen. It has also been shown that metabolites in urine samples can be stable for several weeks under storage in a deep freezer (Saude and Sykes, 2007). Sampling techniques are also a critical concern. Intracellular metabolites are usually analytical targets of metabolomics; however, it is well-known that physio­ logical responses of unicellular organisms or eukaryotic microorganisms (e.g., protozoans, bacteria, and fungi) are governed not only by intracellular metabolic fluxes but also by external conditions such as the composition of culture medium and temperature. In this sense, analysis of both intracellular and extracellular metabolites will give a better picture of metabolic differentiation of unicellular organisms. Metabolic fingerprinting of both internal and external metabolomes proved useful for the determination of the mode of toxic action of commercial fungicides (Allen et al., 2003, 2004).

22.3.2  Qualitative and Quantitative Measurements Numerous analytical techniques have been devised to cover a wide range of metabolites. These methodologies can be divided into (1) separations and (2) qualitative and quantitative measurements. Well-resolved peaks and unambiguous assignments are necessary for satisfactory

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metabolomic results. However, this may not be strictly required in some research studies. For example, identification of selected metabolites that have differential responses is sufficient as a biomarker for the diagnosis of a specific disease (Fiehn, 2002). To gain insights into a biosystem and to find specific biomarkers, however, it is important to unambiguously identify and quantify metabolites. In most metabolomic practices, qualitative and quantitative measurements are integrated with chromatographic or electrophoretic separations, but some techniques that do not require separations are also common [e.g., most nuclear magnetic resonance (NMR)-based methods]. Analytical procedures combined with separation are preferred to identify metabolites in complex mixtures. Representative separation technologies include gas chromatography (GC), high-performance liquid chromatography (HPLC), and capillary electrophoresis (CE). Because numerous publications on these methods exist, they are not discussed in detail in this chapter. The most sophisticated instrumentation in metabolomics is dedicated to structural identification and quantitation. Several different methods, including ultraviolet ­visible (UV-VIS), Fourier transform infrared (FT-IR) spectroscopy, mass spectrometry (MS), NMR, and Raman spectroscopy, are used in metabolomics (Cherney et al., 2007; Harrigan et al., 2004; Lenz et al., 2005; Viant et al., 2003; Wilson et al., 2005a,b). Because each method has advantages and disadvantages, a proper method should be carefully selected for suitable uses. MS and NMR are among the most popular techniques because of their capability to elucidate chemical structures. In addition, MS and MNR spectral libraries contain up to 1 million compounds, which make it more feasible to identify metabolites. Excellent reviews of MS and NMR techniques are available (Dunn and Ellis, 2005; Ramautar et al., 2009; Villas-Boas et al., 2005; Wishart, 2008).

22.3.2.1  Mass Spectrometry MS is widely used in metabolomics because it can provide sensitive, rapid, and qualitative and quantitative analyses of metabolites (Dunn and Ellis, 2005; Ramautar et al., 2009). MS can be operated alone or coupled with a separation technique. Common ionization modes and mass analyzers are summarized in Table 22.2. Several different ionization modes are used in mass spectrometry, depending on the analytes or desired information. The most common ionization modes include electron impact, chemical ionization, electrospray (ESI), and matrix-assisted laser desorption ionization. After ionization, mass analyzers are introduced to separate specific mass ions (i.e., ions with specific mass/charge ratios). There are numerous variants of mass analyzers (see Table 22.2). Quadrupole (Quad), time-offlight (TOF), and ion trap (IT) are the most commonly used mass analyzers. In addition, several tandem mass analyzers

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Table 22.2  Ionization Methods and Mass Analyzers Commonly Used in Metabolomics Studies Class

Name

Description

Mass analyzer

Quadrupole (Q)

Common, reasonable mass resolution

Ion trap (IT)

Common, benchtop, from poor to reasonable mass resolution, MSn determination available

Time-of-flight (TOF)

Common, benchtop, reasonable mass resolution, wide range of linear dynamic response, good precision

Ion cyclotron resonance (ICR)

Rapid and highest resolution, sensitive, limited linear dynamic range, high mass accuracy

OrbiTrap

High mass accuracy, limited linear dynamic range

Electron impact (EI)

Most common, GC compatible, limited to small molecules, largest set of spectral library available

Chemical ionization (CI)

Soft ionization, GC compatible, usually limited to small molecules, higher probability to get molecular ions

Electrospray ionization (ESI)

Soft ionization, common, LC compatible, extensive use in most biomolecules (small metabolites, proteins, etc.)

Matrix-assisted laser desorption ionization (MALDI)

Soft ionization, common, LC compatible, extensive use in most biomolecules (small metabolites, proteins, etc.)

Desorption electrospray ionization (DESI)

Soft ionization, no sample treatment required, raw samples (e.g., tissues, cells), initial optimization required

Extractive ESI (EESI)

Soft ionization, variant of ESI, minimum sample pretreatment, reduced source contamination

Ionization mode

are available (e.g., triple Quad, TOF-TOF, TOF-Q, and IT-TOF). A tandem mass analyzer is particularly useful for analyses of complex mixtures because specific ions in overlapped peaks can be analyzed to specify the molecular identities (Ceglarek et al., 2009; Dettmer et al., 2007). Metabolomics samples can be analyzed by MS techniques with or without prior chromatographic or electrophoretic separations. In practice, direct injection mass spectrometry (DIMS) represents MS techniques without prior separation. Crude extracts are directly injected (or infused) into an MS. Chemical composition and other information can be recovered with complex-featured spectra (specific m/z fragments and their intensity). The best examples of DIMS are Fourier transform-ion cyclotron resonance mass spectrometry (FT-ICR MS) and TOF MS. DIMS has been applied successfully in several studies, and excellent reviews are available (Allen et al., 2003; Brown et al., 2003; Hasegawa et al., 2007; Madalinski et al., 2008; Ohta et al., 2007; Taylor et al., 2009). These mass analyzers are frequently coupled with LC or GC (Baidoo et al., 2008). MS combined with chromatography (LC or GC) or electrophoresis (CE) is widely applied in metabol­omic studies. Because of the rapid advancement of HPLC instrumentation since the 1970s, this technique has been most widely used to analyze numerous metabolites in biological samples. HPLC-MS, also referred to as LC-MS, is

one of the most popular analytical tools in metabolomics. Theoretically, all metabolites, including macromolecules, can be analyzed with this versatile instrument because volatility, which is required in GC, is circumvented. The development of new chromatographic materials and columns (e.g., nano-sized packing materials, monolithic columns, and nano and capillary columns) has resulted in improved resolution power. Among recently introduced LC techniques, hydrophilic interaction liquid chromatography is a noticeable achievement for the separation of highly polar metabolites (Kamleh et al., 2008). Comprehensive metabolomic studies with pesticides and xenobiotics have several advantages compared to conventional research methods. For example, both biochemical effects of a specific toxicant and its metabolism can be analyzed simultaneously through a metabolomic approach. HPLC-based metabolomics is one of the best choices for this purpose because metabolites of pesticides or xenobiotics usually contain reactive functional groups (e.g., hydroxyl, amino, and carboxylic acids) and are not suitable for GC-MS analyses. Most LC-MS techniques use soft ionization modes such as ESI for obtaining molecular information (e.g., molecular weight and formula). A number of excellent metabolomic studies have been performed with LC-MS (Bruder et al., 2004; Buscher et al., 2009; Cheng et al., 2008; Plumb et al., 2002; Weckwerth, 2003).

Chapter | 22  Metabolomics in Pesticide Toxicology

GC-MS is also an important analytical method in metabolomics (Fiehn et al., 2000; Pasikanti et al., 2008). The strength of GC over HPLC is its much higher resolution and convenient hyphenation with MS. A long GC column (30–100 m) can easily resolve more than several hundred components in a crude extract, which is hardly achievable with conventional HPLC methods. Another advantage of GC-MS methods is that many preconstructed spectral databases are available (e.g., NIST and WILEY). Knowledge of spectral patterns is the most important prerequisite for structural identification of metabolites. However, metabolites have to be evaporated for introduction into a GC column. Hence, proper derivatization is frequently required in GC-MS analyses. Common derivatization methods include alkoxime formation, silylation, alkylation, and esterification. Although most small molecular metabolites can be analyzed after these procedures, some metabolites (e.g., phospholipids) are still not amenable for GC-MS, and complementary tools such as LC-MS are required. These additional procedures increase the analytical effort and bias from incomplete or excessive derivatization and may result in the production of multiple peaks from a single compound. Kanani and Klapa (2007) suggested several precautions and correction strategies for GC-MS–based metabolomic data. Although GC-MS–based methods are not an ideal choice for highthroughput analysis, the previously mentioned advantages make GC-MS one of the most feasible tools in metabol­ omics. Applications of GC-MS in metabolomics can be found in disease diagnosis, biomarker discovery, and xenobiotics metabolism (Ippolito et al., 2005; Keum et al., 2008; Kuhara, 2007; Qiu et al., 2008; Remer et al., 2005). CE (or CE-MS), although only recently introduced for metabolomics, has been commonly used in the analysis of highly polar or charged chemicals. Reverse-phase HPLC generally gives good performance for metabolites of moderate to low polarity. However, separation of highly polar or charged metabolites (e.g., phosphorylated sugars and nucleotides) with HPLC is problematic. CE-MS is the ideal method for the resolution and structural confirmation of these cumbersome metabolites. Ramautar et al. (2009) provide excellent examples of CE-MS use on biofluid metabolomics.

22.3.2.2  Nuclear Magnetic Resonance Spectroscopy Since the 1980s, various NMR techniques have been used for metabolic fingerprinting (Nicholson et al., 1985). Advantages for metabolomics research include (1) rapid data acquisition with reduced sample preparation times and (2) high reproducibility. Both factors make NMR a powerful tool for high-throughput applications in metabolomics (e.g., disease diagnosis and phenotyping). In contrast to chromatography-coupled MS, NMR suffers from low

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sensitivity and less comprehensive spectral libraries. Because of these limitations, it is difficult to obtain quantitative and qualitative information of less abundant metabol­ ites with ordinary NMR methods. Another complexity of NMR techniques is related to the fact that multiple signal points are obtained from a single compound, resulting in the complicated spectra obtained from complex mixtures (a common feature of metabolomics samples), which are difficult to interpret. Rapid advances in spectral processing tools, higher field magnets, and other technologies have overcome these difficulties to a large degree, making NMR the most popular metabolomic platform. To date, 1H-NMR is the most widely used method in metabolomics (Lindon and Nicholson, 2008a,b; Viant, 2003; Wishart, 2008). Griffin (2003) noted that given the small chemical shift range of 1H (generally 10 ppm), there is significant overlap between metabolites in one-dimensional NMR. Several approaches have been developed to solve these problems, including multidimensional techniques. Numerous examples are available, including correlation spectroscopy (COSY), HeteroCOSY, and total COSY (Viant et al., 2003; Xi et al., 2007). Novel spin-relaxation sequences are now commonly used in metabolomics research. Several heteroatoms other than 1H (e.g., 13C, 31P, and 15N) can be used for metabolomics either alone or in multi­dimensional NMR analyses (e.g., HeteroCOSY experiments). Because of very low natural abundances of 1.1 and 0.4% for 13C and 15N, respectively, NMR methods that use these nuclei suffer from low sensitivity. However, when NMR is combined with specific isotope-enrichment techniques, valuable biochemical information can be obtained. Kikuchi et al. (2004) successfully utilized isotope labeling techniques (13C and 15N) to determine the relative metabolic flux of low-molecular metabolites in Arabidopsis thaliana. This approach provides not only pathway information but also fluxomic data (Lundberg and Lundquist, 2004). Birkemeyer et al. (2005) provide a review of isotope labeling in metabolomics. Because less pretreatment or purification has been applied in NMR sample preparations, excessive solvent and macromolecules (e.g., water, proteins, and DNAs) can prevent the proper acquisition of signals of low-molecular-weight metabolites. A number of NMR techniques (e.g., a Carr– Purcell–Meiboom–Gill pulse sequence, relaxation editing, and 1H,1H J-resolved spectroscopy) have been shown to enhance metabolite analysis (Dunn and Ellis, 2005; Tang et al., 2004; Van et al., 2003; Viant et al., 2003; Wang et al., 2003). Signals from excess water, which is common in most biological samples, can easily be suppressed by different techniques (e.g., most commonly, presaturation of water signal). One of the most noticeable applications of NMR is metabolomics with solid samples such as organs and cells. Magic angle spinning (MAS)-NMR techniques are becoming common for solid-state sample analyses.

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22.3.2.3  Deconvolution and Structure Identification Instrumental analysis is followed by data processing, which includes spectral deconvolution, peak alignment, and structural identification of specific metabolites. Extensive overlap of signals (e.g., peaks) from different ­ metabol­ites is common both in NMR and in chromatographycoupled MS. Deconvolution, defined as isolation of accurate spectra of specific metabolites from mixtures, is an important aspect in structural identification. Several free and commercial tools are available as stand-alone software or as modular units for commercial instrumentation (e.g., AMDIS and MarkerLynx from NIST and Waters, respectively). Another important issue regarding instrumental analysis is spectral and peak alignment of a large number of spectra (or chromatograms). Because a range of retention time shifts (or chemical shift) of specific metabolites is commonly observed in complex mixtures, appropriate alignment protocols should be applied to accurately compare multiple data. Numerous software and Internet resources are available for accurate alignment and further processing of instrumental analysis data (e.g., MetAlign and SpectConnect). In addition, internal standards (e.g., ribitol, n-alkanes, and sodium 3-trimethylsilyl-2,2,3,3-d4propionate) are frequently used to facilitate the alignment. Although most metabolites can be recovered by combinations of extraction methods, only a portion of the metabolites in a sample can be unambiguously identified. For example, one can easily obtain 200–400 peaks on GC-MS from an ordinary sample; however, only several tens of peaks can be identified from a library search. This discrepancy occurs because most structural identification tools are based on predefined databases in which a limited number of spectra of metabolites are recorded. Because of these difficulties, construction of a comprehensive spectral database is becoming one of the most important tasks in modern metabolomics. Interlaboratory, national, and international collaboration is required to achieve such a goal to cover a wide variety of metabolites. During cooperative efforts to construct a large database, several additional needs have been identified. For example, a common format of instrumental analyses data and spectral library is extremely important for efficient communication between research groups. Several Internet resources dedicated to metabolite spectral libraries and integrated biochemical information are available (Table 22.3).

22.3.3  Data Mining Methodologies In comparison with classical approaches investigating a limited number of biochemical reactions, metabol­ omics is characterized by the simultaneous production of a huge amount of data from a single experiment. In turn, large replicates are common in most omics studies.

Table 22.3  Databases of Metabolomics and Biochemical Pathways Name

Website

Information

Golm Metabolome Database

http://csbdb.mpimpgolm.mpg.de/csbdb/ gmd/gmd.html

Various metabolites with mass spectra

Human Metabolome Database

http://www.hmdb.ca

Metabolites found in human body (6500 entries)

Platform for RIKEN http://prime.psc. riken.jp

Various metabolites with mass spectra

Metabolomics (PRIMe)

NMR

KEGG PATHWAY Database

http://www.genome. Integrated database jp/kegg/pathway.html of biochemical pathways, genes, proteins, and metabolites (no spectral information)

CytoScape

http://www. cytoscape.org/index. php

Data analysis tools for systems biology

Metabolomic samples usually contain extremely diverse metabolites with large variations of relative abundance, which require a large dynamic range of instrument detection. Because of these complexities, extensive statistical treatment is required for metabolomic data processing. In addition, complementary information from other omics is often required for proper interpretation of metabolomics data because the metabolomic differentiation is associated with responses at protein (proteomic) and mRNA (transcriptomic) levels. Important basic statistical issues in metabolomics include detection of outliers and normalization and replacement of empty data when necessary. Classical statistic tools are indispensable for these purposes. Scale reduction of quantitative data is frequently needed because the amount of some metabolites far exceeds that of others (e.g., sugars in storage organs of plants). Logarithmic transformation is typically used to adjust for these differences. Several different normalization methods are suggested to enhance statistical comparisons (e.g., mean, median, and maximum likelihood methods) (Sysi-Aho et al., 2007). A wide variability in individual biosystems (samples) makes it difficult to detect outliers from a complex data set. The most common method is comparing the deviation between mean and median values. Analysis of variance and its variants are used to discriminate statistical differences between different metabolomes.

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Another common statistic used in omics is related to data reduction (dimension reduction) procedures. Because of the huge amount of data, it is difficult to extract specific patterns from omics data through manual inspection, so several data reduction methods are generally used. These include principal component analysis (PCA), discriminant analysis (DA), partial least squares (PLS), and other recently introduced methods (e.g., support vector machine and neural network) (Beckonert et al., 2003; Bijlsma et al., 2006; Goodacre et al., 2004; Lindon and Nicholson, 2008b; Qiu et al., 2008; Rubingh et al., 2006; Wang et al., 2003). Although mathematical algorithms and details are quite different, complex metabolomic data can be described with a reduced number of variables (i.e., components or discriminants) through these statistical treatments. In addition to data dimension reduction, integrated approaches with other omics databases are becoming common for verification and visualization of metabolomic differentiations. For example, one can easily locate and visualize the differential response of metabolomes in common biochemical pathways using KEGG Atlas (provided by Kyoto Encyclopedia of Genes and Genomes) (see Table 22.3). Okuda et al. (2008) reported the potential utility of KEGG Atlas for global analysis of metabolic pathways and metabol­ omics. For the evaluation of connectivity between metabol­ ites and other omics data, CytoScape has been developed through the cooperative efforts of many systems biologists.

22.3.4  Emerging Technologies With Novel Applications Advances in metabolomics have occurred both in instrumentation and in theoretical knowledge. In relation to technological aspects, numerous novel tools have been introduced. Ordinary experimental protocols routinely include sample extraction procedures, where the efficiency may be problematic. To circumvent potential problems, several methods requiring in situ or minimum pretreatment have been developed. Although these methods should be rigorously confirmed for routine analyses, their applications in metabolomic research have been reported (Cha et al., 2008; Dettmer et al., 2007, Gu et al., 2007; Weckwerth et al., 2004). Several new separation tools are becoming common in metabolomics, including multidimensional chromatography and selective enrichment techniques (Carlson and Cravatt, 2007a,b; Huang and Regnier, 2008; Koek et al., 2008). Multidimensional techniques such as two-dimensional (2D) GC and 2D-HPLC can resolve more metabolites in a complex mixture with increased sample loading for each instrumental run. In addition, several chromatographic materials with improved resolution performance, such as monolithic columns in HPLC, are now available. The recent development of ultraperformance liquid chromatography

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(UPLC) has resulted in a greater sensitivity, enhanced resolution, and shorter analysis time than those of traditional HPLC. UPLC provides more than double peak capacity, an approximately 10-fold increase in speed, and a 3- to 5-fold increase in sensitivity compared to conventional reversephase HPLC (Wilson et al., 2005b). UPLC coupled with MS (UPLC-MS) has been shown to improve metabolic classification or differentiation under chemical and physical stimuli (Michopoulos et al., 2009; Wilson et al., 2005b; Zelena et al., 2009). Tandem LC-MS and CE-MS such as triple Quad, TOF-Q, and IT-TOF will continue to play an important role in metabolomics research. Novel applications of soft ionization modes such as desorption ionization have been applied in several metabolomics areas. The most interesting advantage of the emerging desorption ionization techniques is that little sample pretreatment is required and metabolites in solids can be monitored. Cha et al. (2008) applied a colloidal graphite laser-desorption ionization technique to localize several metabolites in intact Arabidopsis leaves. Additional novel applications in this area have been reported (Bedair and Sumner, 2008; Gu et al., 2007). Some NMR methods, such as MAS-NMR, routinely used for chemical identification in solid state are now being applied to the analysis of intact tissues or cells (Yamamoto et al., 2007). Because of the line-broadening effects, it is difficult to obtain clear spectra from solid samples with classical NMR instruments. MAS spectroscopy, however, provides a path to circumvent this limitation (Shockcor and Holmes, 2002). MAS-NMR aptly serves as a tool for linking biofluid changes to mechanism of action in target tissues (Lindon and Nicholson, 2008a,b; Robertson, 2005). There are several examples of the use of MAS technology integrated with traditional NMR-based metabolomics (Coen et al., 2003, 2004; Garrod et al., 2001; Southam et al., 2008). In practice, this technique requires specialized rotors and magnets and high-throughput analysis, and it may not be applicable in the near future. With rapid advances in analytical instrumentation, combinations of tools are becoming available. Most notable is the hyphenated NMR approach (with LC-NMR or LC-NMR-MS) (Clarkson et al., 2005; Deighton, 2008). Other useful but not frequently used methods include vibrational spectroscopy (IR, near-IR, and Raman spectroscopies). These methods are limited in their ability to resolve metabolites in complex mixtures. Some metabol­ ites with characteristic vibrational spectra, however, have been easily characterized and used as markers (Baranska and Schulz, 2005; Cherney et al., 2007; Harrigan et al., 2004). These spectroscopy techniques are nondestructive and thus can be used for in situ analyses but not in tandem with MS. With the increased complexity and size of data, efficient processing of analytical data is becoming more important than any other aspect of metabolomics. When we consider

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the final goal, metabolomic results should be interpreted in an integrated manner with genomic and proteomic information available. Several chemometric databases (e.g., MS and NMR spectral libraries) are now integrated with other information sources, particularly bioinformatics, and efforts to expand the coverage are ongoing.

Table 22.4  Representative Applications of Metabolomics Application area

Purpose

Description

Chemical toxicology

Determination of toxic effects on primary and secondary metabolism

Qualitative and quantitative profiles of whole metabolites

Diagnostic chemistry

Diagnosis of disease and metabolic malfunction Phenotyping of metabolic disorders

Biomarker development and application by metabolic fingerprints

Drug and pesticide discovery

Discovery of novel targets (enzymes and metabolisms)

Qualitative and quantitative profiles of whole metabolites, metabolic correlation between other metabolites, genes, and proteins

Plant biology

Phenotyping, discovery of novel metabolic pathways and genes

Phenotyping of specific plants by metabolic fingerprints Speciation of metabolic pathways of secondary metabolites

22.4  Applications It is well-known that metabolites in primary metabolism are highly conserved through simple microorganisms to mammals, whereas gene sequences of specific enzymes are highly diverse. Furthermore, there can be several different sets of functionally equivalent enzymes with diff­ erent amino acid sequences in the same organism. In this sense, the metabolome is a more common biological language than the genome or proteome. These simplicities make it more feasible to apply metabolomics for various scientific interests. In addition, metabolomics studies can give better insights into physiological effects of specific stimuli because changes in proteins or genes do not always result in phenotypic differentiation. Major applications of metabolomics are found in biomarker discovery, phenotyping, disease diagnosis, and risk assessment of genetically modified organisms (GMOs). Related research articles can be found in pharmacotoxicology, environmental sciences, nutrigenomics, plant biology, and many others. The following discussion focuses on applications of metabolomics in pesticide toxicology and environmental sciences. Toxicological response is one of the most common subjects of current metabolomics studies. Toxicology is a discipline of study of adverse effects of chemicals on living organisms, which includes biological response to toxicants. The term “response” refers to physiological, biochemical, and morphological changes stimulated by stressors whether they are chemical, physical, or biological. Toxicological studies via metabolomic approaches are comparative and comprehensive. Analytical data represent qualitative and/or quantitative information (e.g., identity and concentrations of metabolites) in test samples. The data require further data mining, analysis, and synthesis to understand the biological relevance. Metabolic fingerprinting refers to the comparison of data (signals) patterns (not necessarily exact identities of metabolites) of complex chromatograms or spectra with those in a database. Metabolomics research in diagnostics and taxonomy can be classified in this category, in which specific patterns alone are adequate to meet the analytical purpose. Comprehensive profiling (data collection) methods focus on both quantitative and qualitative information of entire metabolites. Similar analyses for a predefined set of metabolites in specific biochemical pathways are designated as “targeted metabolomics”. Major applications of these methods are summarized in Table 22.4.

Among the representative applications, the largest proportion of current research is attributed to clinical toxicology, specifically diagnostic marker development and its use. Because of its robustness and short sample preparation times, NMR is the most popular technique in metabolic fingerprinting of specific diseases, congenital disorders, and cancers (Griffin, 2003; Griffin and Kauppinen, 2006; Griffin et al., 2004; Qiu et al., 2008; Shockcor and Holmes, 2002; Sreekumar et al., 2009; Xu et al., 2009). A criticism of metabolomics as a diagnostic tool is that many metabolomic analyses of biofluids (urine and serum) commonly give “usual suspects” as biomarkers (Robertson, 2005). Differential accumulations of some organic acids and creatine are examples of these so-called usual suspects. The quantitative changes of these metabolites are common in stressed biological organisms, regardless of the nature of the stimuli (Robertson, 2005). Metabolomics data should be understood on the basis of pattern changes rather than simple up-/downregulation of specific metabolites. Although detailed chemical information of constituents is rarely provided in metabolic pattern recognition research, peak profiles in chromatograms or spectra are sufficient to diagnose some diseases. Comprehensive and detailed information of cellular responses can be obtained when fingerprints are accompanied by structural information.

Chapter | 22  Metabolomics in Pesticide Toxicology

Sreekumar et al. (2009) found that 87 metabolites were differentially accumulated in cancer cells during the progression of the disease, and structural analyses allowed for identification of sarcosine as a potential agent in cancer progression. Excellent reviews by Griffin and Shockcor (2004) and Griffin and Kauppinen (2006) describe several additional examples. Pesticide discovery requires a multitiered approach that includes screening for bioactivity, adverse effects (toxi­ city), dose–response (efficacy), and metabolism (detoxification). In addition to a parent pesticide, in vivo toxicity assessment of the pesticide becomes more complicated because the extensive metabolism of chemicals can alter the potency or toxicological profiles. Another important issue in chemical toxicology is the biochemical effects at marginal or sublethal concentrations. When the timedependent response is of concern, one must consider additional factors such as acute, subacute, and chronic toxicity. Quite different physiological changes can be observed over time. Elucidation of mode of action is critical in pesticide and drug discoveries. However, it can be a laborious and time-consuming task because a biosystem contains a wide array of metabolites, proteins, and genes. Metabolomic approaches can provide high-throughput analytical options for this purpose. Clearly separated metabolic patterns were observed in Saccharomyces cerevisiae treated with 10 commercial fungicides, from which two large clusters were derived according to the inhibitory activities of respiration (Allen et al., 2004). Ott et al. (2003) utilized NMRbased metabolic fingerprinting techniques to classify 27 herbicides according to the mode of action. They could successfully distinguish between pesticides with different modes of action. Oikawa et al. (2006) suggested differential effects of several herbicides, based on FT-ICR MSbased analyses, combined with a metabolic relationship database (KNApSAK). Although most mode-of-actionrelated metabolomics studies have been performed with NMR, a recent application of GC-MS for herbicide metabolomics to specify the mode of action revealed that separation-based metabolomics may be more informative by using detailed analysis of metabolomes (Trenkamp et al., 2009). For the identification of mode of action of novel chemicals, Aliferis and Chrysayi-Tokousbalides (2005) used a comparative approach to compare the spectral fingerprints of plant extracts from novel chemical treatments with those from herbicides with known mode of action and NMR data. They suggested that the mode of action of pyrenophorol differs from that of commercial herbicides, and the natural product can be a novel template for new pesticide development. Metabolomics has been used in drug discovery in several ways. Possible targets of drugs can be identified through the comparative analysis of metabolic patterns of normal and diseased samples. Differential metabolomic

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responses to closely related drugs can be useful information for structural modification to improve the properties of specific drugs (Kaddurah-Daouk et al., 2008). By a general definition, target analytes in most metabolomics are naturally produced metabolites. However, Chen et al. (2007) suggested a promising utility of modern metabolomic technologies to monitor xenobiotics metabolism. Because drugs (or pesticides) and their metabolites may have different potencies, fates, and effects on primary metabolism, all of which can induce metabolic responses, a comprehensive profiling of xenobiotics and metabolites is as important as or more important than that of metabolomics of primary metabolism. Aubert and Pallett (2000) successfully used 13 C- and 19F-NMR to investigate the mode of action and metabolism of the herbicide isoxaflutole. Regarding the widespread use and public concerns, evaluations of pesticide toxicity with omics tools are far more limited than those of human drugs or diseases. Similarly, few publications concerning sublethal effects of organic pollutants from a metabolomics perspective are available. However, related research has gained some attention from scientific communities and regulatory agencies. Bundy et al. (2001) proved the usefulness of NMRbased metabolomics for speculating on the toxic responses of soil invertebrates under chemical stress. Since then, many articles related to pesticide toxicology have been published. As mentioned previously, in vivo evaluation of pesticides is a quite complicated issue, especially in higher animals. A pesticide can follow diverse metabolic pathways in different organisms. Many metabolites with different toxicological and physicochemical properties are often produced after exposure to a specific pesticide (Coleman et al., 2000; Hodgson, 2001; see Chapter 38). In general, recently registered pesticides show improvements in performance compared to some legacy pesticides such as organochlorine insecticides. These new pesticides show increased selectivity, shorter half-lives, and a decrease in side effects, although some unintentional effects (e.g., toxicity to nontarget organism and harmful effects other than major mode of action) are inevitable. Most pesticides have a specific known mode of action. For example, organophosphorus and N-methyl carbamate insecticides are metabolic inhibitors, although the final outcome is the inhibition of neuronal signal transfer, specifically on acetylcholine catabolism. Azole fungicides are well-known inhibitors of cytochrome P450, which is responsible for fungal ergosterol biosynthesis. Some organochlorine insecticides modulate the ion mobility in biological membranes. Some fungicides, such as fenpiclonil and fludioxonil, act by intervening in signaling pathways of osmotic stress through the inhibition of nonmetabolic sensory proteins. Considering the wide variety of mode of action, it is quite surprising that some metabolites are commonly found in pesticide-stressed animals. These so-called

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usual suspects include creatine, choline, and some osmolytes, nucleotides, lactate and related short-chain organic acids, and alanine. Ekman et al. (2006) analyzed metabolomic differences between normal and triazole fungicide-treated rats. NMR spectra of polar extracts from several organs (liver and testis) and sera were obtained and evaluated with PCA and PLS-DA. After rigorous statistical analysis, differential accumulation and depletion of some of the previously mentioned metabolites in a dose-dependent manner was observed. Clearly distinguishable metabolomic patterns were also observed between the treatments of different fungicides and those of control. Some researchers at the U.S. Environmental Protection Agency performed 1D- and 2D-NMR metabolomics to examine the toxicity of vinclozolin in fathead minnows (Pimephales promelas), a model organism of aquatic toxicity testing (Ekman et al., 2007). They confirmed that levels of several unidentified metabolites were up- and downregulated in combination with the differential patterns of usual suspects. Viant et al. (2006a) determined metabolomic responses to dinoseb, diazinon, and esfenvalerate in Chinook salmon (Onchorhynchuss tshawytscha). Notable differences were found particularly in ATP and some amino acids, depending on pesticides and concentrations. Some marker metabolites (vitellogenin and cholesterol) were found from comprehensive metabol­ omics of rainbow trout (Oncorhynchus mykiss) exposed to the synthetic contraceptive estrogen ethinylestradiol (Samuelsson et al., 2006). One-dimensional 1H-NMR is the most common technique in chemical toxicologyrelated metabolomics. However, applications of 13C-NMR in combination with modified sample extraction have demonstrated that heteronuclear NMR is another promising tool (Jahns et al., 2008). In toxicological studies with mice, Jahns et al. (2008) applied 13C-NMR in a metabol­ omics study and found an extensive redistribution of lipids (fatty acids, triacylglycerols, and steroids) after exposure to TCDD. Although nuclei other than 1H and 13C are not commonly used in NMR-based metabolomics, Viant et al. (2006b) successfully used 31P-NMR to quantitatively measure differences of phosphorylated metabolites in dinosebtreated Japanese medaka (Oryzias latipes). Toxicological metabolomics has been performed with several invertebrates including earthworms (mostly Eisenia spp.) exposed to toxic stressors (Bundy et al., 2001, 2002a,b, 2008). Metabolomic responses of several earthworm species (Eisenia spp. and Lumbricus rubellus) have been evaluated after exposure to herbicides, insecticides, and other environmental pollutants (Guo et al., 2009; McKelvie et al., 2009). In a multiplatform analysis of earthworm metabolomes, McKelvie et al. (2009) found that levels of sugars and amino acids (e.g., maltose, glycine, alanine, and leucine) were affected by DDT and endosulfan. The results of comparative analyses suggested the ratio of alanine to glycine as a potential biomarker of

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pesticide exposure. Rochfort et al. (2009) observed a clear indication that different metabolomic patterns of earthworms are relevant to soil health. Physiological responses of blue mussel (Mytilus edulis) to several physical and chemical stressors were evaluated with 1H NMR-based metabolomics (Tuffnail et al., 2009). In this case, alanine and some osmolytes were the important metabolites that helped in the discrimination among different stress types. Toxicological effects of atrazine and heavy metals on crustaceans were tested using MS-based metabolomics (Ralston-Hooper et al., 2008; Taylor et al., 2009). From these studies, several amino acids and amines were characterized as useful biomarkers through the analysis of metabolomes by GCxGC-TOF and FT-ICR MS. With the rapid advancement of genetic technologies, vigorous debates arise over the safety of GMOs. The subject of biochemical equivalence between transgenic and wild-type organisms is quite contentious and, unfortunately, metabolomic applications to address this problem have been rather limited. A hierarchical metabolomics study demonstrated substantial compositional similarity between genetically modified and conventional potato crops (Catchpole et al., 2005). GC TOF-MS and flow injection ESI-MS were used to cover a wide variety of metabolites, such as simple organic compounds, polysaccharides, and alkaloids. Zhou et al. (2009) investigated the metabolic equivalence of transgenic rice to wild-type rice regarding several genes related to the biosynthesis of the insecticidal Bt toxin and trypsin inhibitors and found nonequivalent profiles for some metabolites. Stamova et al. (2009) observed metabolic profiling of transgenic wheat that overexpressed the high-molecular-weight Dx5 glutenin subunit after MS-based metabolomics. It is expected that research on metabolite profiles in GM crops will continue to increase in response to public concerns and academic interests. Metabolomic analysis alone, however, cannot provide sufficient safety-related information. Comprehensive assessments should be applied using data from integrated omics studies and toxicological responses.

Conclusions Metabolomics research has provided valuable insights into many scientific interests. As a result of rapid technological advances, many more metabolomics results will abound in the near future. However, metabolomics has weaknesses (Miller, 2007). Because toxicological responses are under the regulation of a complex array of genes, proteins, and metabolites, combinations of different omics approaches are required to understand biosystems. There are several examples of multi-omics work, and these studies will be more common in the future (Coen et al., 2004; Lindon and Nicholson, 2008a; van Brummelen et al., 2009). These trends will improve both the quality and the quantity of

Chapter | 22  Metabolomics in Pesticide Toxicology

metabolomic data, especially in chemical toxicology. A combinatorial approach of metabolomics with proteomics and genomics, for example, offers great potential to facilitate the cleanup of pesticide-contaminated sites (Singh, 2006). Most of the current research in these areas remains either diagnostic in nature or is limited to fingerprinting. Interpretation of metabolomic data integrated with other omics data will provide truly systems biology-like information. Another important task is standardization of general protocols, including experiments, data handling, and reporting. Use of a common form of experimental description and data reporting can enhance efficient comparison of metabolomic data as well as exchange with other scientific communities. The Metabolomic Standards Initiative (MSI) was launched recently, bringing that goal closer to fruition (Fiehn et al., 2007). Useful suggestions from MSI are available (Goodacre et al., 2007; Morrison et al., 2007). Finally, it should be noted that only a limited number of organisms have been evaluated using metabol­omics in relation to toxicometabolomics. No metabol­omicsrelated publications are available for avian species, insects, amphibians, and so on. Metabolomics research that is focused on these organisms will provide a more comprehensive picture of the ecological effects of pesticides.

Acknowledgments This work was supported in part by grants from the State of Hawaii, the Department of Agriculture and the U.S. Fish and Wildlife Service, USDA Tropical and Subtropical Agricultural Research awards, U.S. EPA award 98951201-1, U.S. NRL award N00173-05-2-C003, and Hawaii Energy and Environmental Technologies Initiative Award N00014-09-1-0709. We thank Margaret R. Ruzicka at the University of Hawaii–Manoa and Maria S. Sepúlveda at Purdue University for their comprehensive review of this chapter.

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Wang, Y., Bollard, M. E., Keun, H., Antti, H., Beckonert, O., Ebbels, T. M., Lindon, J. C., Holmes, E., Tang, H., and Nicholson, J. K. (2003). Spectral editing and pattern recognition methods applied to highresolution magic-angle spinning 1H nuclear magnetic resonance spectroscopy of liver tissues. Anal. Biochem. 323, 26–32. Weckwerth, W. (2003). Metabolomics in system biology. Annu. Rev. Plant Biol. 54, 669–689. Weckwerth, W., Wenzel, K., and Fiehn, O. (2004). Process for the integrated extraction, identification and quantification of metabolites, proteins and RNA to reveal their co-regulation in biochemical networks. Proteomics 4, 78–83. Wiechert, W., Schweissgut, O., Takanaga, H., and Frommer, W. B. (2007). Fluxomics: mass spectrometry versus quantitative imaging. Curr. Opin. Plant Biol. 10, 323–330. Wilson, I. D., Plumb, R., Granger, J., Major, H., Williams, R., and Lenz, E. M. (2005a). HPLC-MS-based methods for the study of metabon­ omics. J. Chromatogr. B 817, 67–76. Wilson, I. D., Nicholson, J. K., Castro-Perez, J., Granger, J. H., Johnson, K. A., Smith, B. W., and Plumb, R. S. (2005b). High resolution “ultra performance” liquid chromatography coupled to oa-TOF mass spectrometry as a tool for differential metabolic pathway profiling in functional genomic studies. J. Proteome Res. 4, 591–598. Wishart, D. S. (2008). Quantitative metabolomics using NMR. Trends Anal. Chem. 27, 228–237.

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Xi, Y., deRopp, J. S., Viant, M. R., Woodruff, D. L., and Yu, P. (2007). Automated screening for metabolites in complex mixtures using 2D COSY NMR spectroscopy. Metabolomics 2, 221–233. Xu, J., Zhang, J., Cai, S., Dong, J., Yang, J. Y., and Chen, Z. (2009). Metabonomics studies of intact hepatic and renal cortical tissues from diabetic db/db mice using high-resolution magic-angle spinning 1 H NMR spectroscopy. Anal. Bioanal. Chem. 393, 1657–1668. Yamamoto, T., Horii, I., and Yoshida, T. (2007). Integrated NMR-based metabonomic investigation of early metabolic effects of ethylene glycol monomethyl ether (EGME) on male reproductive organs in rats. J. Toxicol. Sci. 32, 515–528. Zelena, E., Dunn, W. B., Broadhurst, D., Francis-McIntyre, S., Carroll, K. M., Begley, P., O’Hagan, S., Knowles, J. D., Halsall, A., Wilson, I. D., and Kell, D. B. (2009). Development of a robust and repeatable UPLC-MS method for the long-term metabolomic study of human serum. Anal. Chem. 81, 1357–1364. Zhan, X., and Desiderio, D. M. (2003). A reference map of a human pituitary adenoma proteome. Proteomics 3(5), 699–713. Zhou, J., Ma, C., Xu, H., Yuan, K., Lu, X., Zhu, Z., Wu, Y., and Xu, G. (2009). Metabolic profiling of transgenic rice with cryIAc and sck genes: an evaluation of unintended effects at metabolic level by using GC-FID and GC-MS. J. Chromatogr. B 877, 725–732.

Section IV

Dermatotoxicology of Pesticides

(c) 2011 Elsevier Inc. All Rights Reserved.

Chapter 23

Irritant Dermatitis Lisa E. Maier1, Howard I. Maibach2 and Michael O’Malley3 1

University of Michigan, Ann Arbor, Michigan, and Veterans Administration Medical Center, Ann Arbor, Michigan University of California, San Francisco, California 3 California Environmental Protection Agency, Sacramento, California, and University of California, Davis, California 2

23.1  Introduction

Irritant contact dermatitis (ICD) is defined as nonimmunologic skin inflammation after contact to a substance or physical factor. Although epidemiologic data are scarce, ICD appears to be an important cause of occupational and nonoccupational skin disease. The U.S. Bureau of Labor and Statistics estimated that 80% of occupational contact dermatitis cases were due to ICD in 1995 (Chew and Maibach, 2003). The burden of ICD in agricultural workers is unknown; however, it is likely high given the potential exposures to irritants in agriculture. This chapter discusses the factors influencing irritant potential, delineates general clinical presentations of irritant dermatitis, and addresses workup and treatment. In addition, it addresses methods of evaluating a chemical’s irritant potential and discusses the irritation potential of some agricultural chemicals and plants.

23.2  Factors influencing irritant potential

and balance or result in protein denaturation (Welss et al., 2004). These disruptions compromise skin barrier function, resulting in increased transepidermal water loss and inflammation. Beyond the effect on the stratum corneum, some irritants may directly damage cell membranes and cell proteins. Disruption of cell membranes triggers an inflammatory cascade that results in erythema and edema. Cell membrane damage may also result in abnormal signal transduction (Welss et al., 2004). Irritant potential is also dependent on the molar concentration (Tupker, 2003) and volume and duration of the exposure (Aramaki et al., 2001). As a general rule, increasing concentration and exposure time and frequency will increase irritant potential. Paradoxically, in some cases, repeated exposure results in improvement of the dermatitis. This is known as a “hardening” effect, in which the skin adapts to the topical irritant. The molecular mechanisms are not completely elucidated; however, changes in stratum corneum thickness and function, downregulation of the production of inflammatory mediators, and alteration in the production of stratum corneum lipids have been observed (Watkins and Maibach, 2009; Welfriend and Maibach, 2008).

23.2.1  Chemical Factors Various factors influence a chemical’s irritant potential. Intrinsic molecular properties such as molecular structure, size, ionization state, lipid solubility, and pKa (Berner et al., 1990; Welfriend and Maibach, 2008) determine the chemical’s interaction with the skin barrier and epidermal cells. For example, chemicals such as organic solvents can cause extraction of stratum corneum lipids (Fluhr et al., 2008). Other irritants may alter the lipid composition Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

23.2.2  Physical Factors Physical factors such as extremes of temperature and humidity, as well as mechanical factors such as occlusion and friction, can enhance irritation of chemicals or act as irritants. Several studies evaluating irritancy of surfactants, perfumes, and detergents demonstrated increased irritation with increased temperature (Berardesca et al., 1995; Clarys 647

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et al., 1997; Fluhr et al., 2005; Rothenborg et al., 1977). High temperatures and humidity also promote sweating, which can increase penetration of the irritant chemical in the skin. Furthermore, sweat acts as an irritant if on the skin for prolonged periods (Slodownik et al., 2008). Experimentally, cold and dry weather increases the irritant potential of sodium lauryl sulfate (SLS) and sodium hydroxide on skin exposed to the environment (Agner and Serup, 1989; John and Uter, 2005; Loffler and Happle, 2003). Low humidity alone acts as an irritant, as evidenced by the common wintertime occurrence of asteatotic dermatitis (Robert, 2003). Occlusion possibly increases irritation by increasing percutaneous absorption of the irritant chemical and decreasing passive transepidermal water loss (TEWL) at the site (Van der Valk and Maibach, 1989). One clinical example is the prolonged, repeated use of occlusive gloves, which may promote abnormal barrier function and the development of cumulative irritant dermatitis (Ramsing and Agner, 1996). Lastly, mechanical friction and pressure can damage the skin barrier, resulting in greater irritation (McMullen and Gawkrodger, 2006). Farage (2006) developed a technique to assess the influence of friction on overall irritation potential by applying various products, including fabrics, menstrual pads, and lotion-coated samples, to the popliteal fossa using an elastic knee band. Normal movements in this location create friction at the test site, inducing mechanical irritation. In this study, the addition of mechanical irritation increased overall irritation of the products tested.

23.2.3  Endogenous Patient Characteristics Endogenous factors such as age, anatomical site, preexisting dermatologic conditions, and genetic background may influence an individual’s predisposition to irritant dermatitis. There is a decreased susceptibility of irritation with increasing age, with children younger that 8 years being most susceptible to skin irritation (Robinson, 1999, 2002; Welfriend and Maibach, 2008). The etiology of this difference is unknown, but changes in structural lipids, cell composition, and renewal have been hypothesized (Welfriend and Maibach, 2008). In addition, the anatomical site of irritant exposure may also influence the likelihood of reaction. In a study by Cua et al. (1990), measurements of TEWL were taken after exposure to SLS on various body sites. The most vulnerable site was the thigh, followed by the upper arm, abdomen, upper back, dorsal and volar forearm, postauricular skin, and ankle. The palm was the least affected. Another study employing the technique of corneosurfametry demonstrated the following regional differences: the forehead, back, neck, and dorsal foot were more easily irritated than the dorsal hand and volar forearm (Henry et al., 1997). The reason for these differences is unknown; however, several studies have demonstrated variable skin permeability

Hayes’ Handbook of Pesticide Toxicology

based on location (Cronin and Stoughton, 1962; Feldmann and Maibach, 1967; Wester and Maibach, 1985). It is reasonable to assume this variation in permeability is responsible for differences in susceptibility to irritation; however, a direct correlation between permeability of the skin, skin thickness, and likelihood of irritation has not been demonstrated (Robinson, 2002). Preexisting dermatologic conditions may increase irritation susceptibility. Atopic dermatitis has frequently been cited as a predisposing condition for irritant dermatitis. Basketter et al. (1996) demonstrated increased irritant response to sodium dodecyl sulfate (SDS) in atopics over control subjects. Furthermore, up to 45% of adults who had atopic eczema in childhood develop hand eczema, which in most cases is irritant contact dermatitis (Thestrup-Pedersen, 2000). A constitutionally compromised skin barrier may be responsible for these findings (Thestrup-Pedersen, 2000). Fillagrin is a protein that plays an important role in stratum corneum architecture and function, and when abnormal or decreased it can result in compromise of skin barrier function. Not surprisingly, de Jongh et al. (2008) described an association with loss of function of the fillagrin gene and increased risk of chronic irritant dermatitis. Some atopic patients have defects in fillagrin, thus explaining the reported susceptibility to irritation. Moreover, another condition with a fillagrin mutation, ichthyosis vulgaris, may increase irritant susceptibility (Welfriend and Maibach, 2008). Despite these studies, some refute the association between atopy and irritation (Basketter et al., 1998; Santucci et al., 2003). Further studies should be undertaken with a variety of potential irritants and possibly separating atopic groups by fillagrin mutation classification. Other endogenous factors such as gender and race may influence irritant susceptibility, but multiple studies have had mixed results. No consistent difference has been noted between men and women or between various racial groups (Robinson, 2001, 2002; Welfriend and Maibach, 2008).

23.3  Identifying suspected irritants Although in theory any substance can cause irritation, some substances pose a greater hazard to human skin than others. To produce, transport, and use various chemicals safely, it is important to identify the irritant potential of chemicals. The majority of regulatory authorities rely on data from animal testing to assess irritant potential. Several methods have been described to identify and characterize possible irritants and quantify irritant potential; most commonly used is the Draize rabbit skin test. This and other animal tests are covered in Chapter 28. One obvious criticism of animal assays is the inherent difference between animal and human skin. Some chemicals cause more irritation in rabbits than humans and vice versa (Nixon et al.,

Chapter | 23  Irritant Dermatitis

1975; Phillips et al., 1972). Furthermore, animal studies such as Draize testing do not simulate “real-world” conditions such as cumulative exposure, high temperature, and compromised skin barrier – all factors that can change the irritant potential of a chemical. In addition, in recent years there has been increased concern for animal suffering. This has resulted in a ban on some animal testing for products in Europe (Brekelmans, 2007). Several human-based models have been described to address these issues, and they may serve as future approaches to best obtain this vital data on irritation.

23.3.1  Irritant Patch Testing In 1977, the National Academy of Sciences reported a human single application patch test procedure in which occlusive patches are applied on the intrascapular region of the back or dorsal forearms (National Academy of Sciences, 1977). The duration of exposure is variable depending on the desired study design. Irritation is graded on a visual scale similar to the Draize scale, and it is often compared to response of a reference material as a control. In a variant test, known as the 4-hour patch test (Robinson et al., 2001), 0.2 ml of the test liquid or 0.2 g of solid test material is applied in a Hill Top Chamber containing a Webril pad. Patches are then applied to the upper outer arm of approximately 30 subjects for initially short durations such as 15 or 30 min. Patches are left on for up to 4 h until a positive result occurs. The sites are then graded for degree of visual irritation (e.g., erythema and edema) immediately and 24, 48, and 72 h after patch removal. Irritation is graded as 0, , , or . A grade of  or higher is considered positive. The degree of irritation due to the test chemical is compared to irritation caused by a positive control, 20% SDS. The proportion of individuals with a positive irritant reaction to the test chemical after exposure up to 4 h is measured. If this proportion is significantly less than the proportion of positive irritant reactions to SDS, then the chemical is not classified as an irritant. If there is a similar or higher proportion of irritation, then the chemical is considered an irritant. It is essential to note that these irritant patch testing techniques are used for experimental purposes only. Known irritants should not be patch tested in the clinical setting to confirm irritant dermatitis.

23.3.2  Cumulative Irritation Testing Cumulative exposures to irritants are common; thus, assays to assess the long-term irritation potential of chemicals are important. Several cumulative irritation assays have been described. As described by Lanman et al. (1968) and Phillips et al. (1972), a 1-inch square of Webril is saturated with liquid or 0.5 g of a viscous substance and applied to a

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pad. This pad is then applied to the upper back with occlusive tape. Every 24 h the tape is removed and the test site is examined. This process is repeated for 21 days. Variants of these tests use different durations of exposure such as described by Robinson (2001) and Wigger-Alberti et al. (1997).

23.3.3  Chamber Scarification Test This test is designed to test the irritant potential of products on damaged skin (Frosch and Kligman, 1976). Artificial skin wounding is achieved by superficially scratching six to eight 1-mm sites on the volar forearm. Care is taken not to cause bleeding. Test material in a quantity of 0.1 g is placed in Durhing chambers (for solids) or fitted saturated pads (for liquids). These are placed on the scratched test sites once daily for 3 days. Once pads are removed, irritation is measured via a visual score of erythema and edema. This score can be compared to the product’s effect on intact skin by calculating a “scarification index.” This is the score of the scarified sites divided by the score of the intact sites. It is not known if this test is a reliable model for predicting response of routine use on damaged skin.

23.3.4  Immersion Tests These tests were devised to evaluate real-world use of potential irritants that are often used in “wet work” situations. The term wet work refers to prolonged exposure to liquids, occlusive gloves, hand washing, and water-soluble irritants (Diepgen and Coenraads, 1999). In one model described by Kooyman and Snyder, solutions of soap up to 3% were prepared in basins at 105°F. Subjects immersed one hand and forearm in each basin for 10–15 min, three times a day for 5 days or until irritation occurred (Levin and Maibach, 2008). Evidence of irritation on a visual scale was evaluated.

23.3.5  Bioengineering Approaches Bioengineering tools may allow a more sophisticated and precise assessment of irritation than a simple visual scale. These techniques include assessing for transepidermal water loss, laser Doppler flowmetry, laser Doppler perfusion imaging, capacitance, and chromametric analysis (Bashir and Maibach, 2001). Transepidermal water loss is the water that escapes from the skin surface as a normal process. This measurement has often been used as a method of indirectly assessing barrier function. It is thought that the higher the TEWL, the less effective the barrier. There is some concern about the accuracy and validity of TEWL as a marker for barrier function (Chilcott et al., 2002). Additional studies are needed to assess this

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issue. Laser Doppler flowmetry is a noninvasive technique that utilizes laser light to measure blood flow in the skin (Berardesca et al., 2002). In laser Doppler perfusion imaging, the investigator scans a skin area with a lowenergy laser to measure cutaneous perfusion. The results are projected on a computer screen (Aspres et al., 2003). Electrical skin capacitance is an indirect measurement of the stratum corneum hydration. Other bioengineering techniques exist and are discussed in more detail in other texts, such as Marzulli and Maibach’s Dermatotoxicology (Levin and Maibach, 2008). One other technique that deserves mention is corneosurfametry, which assesses the damage to corneocytes by surfactants. Cyanoacrylate skin surface strippings are harvested from various sites and then exposed to surfactant for 2 h. Subsequently, the stripping is stained with basic fuchsin and toluidine blue and measured by colorimetry. Reflectance colorimetry is used to measure color intensity. The intensity increases with increased irritation. This technique appears to have lower interindividual variability than patch testing and is reproducible (Piérard et al., 1994, 1995).

23.3.6  New Approaches Because of intraspecies disparity, concern for animal welfare, and a desire for more accurate testing methods, the U.S. National Research Council Committee on Toxicity Testing and Assessment of Environmental Agents has issued a statement that new approaches to toxicology testing should be developed. The hope is to move away from animal tests to more testing employing molecular technology, computer modeling, and computational biology (National Research Council, 2007). Some predictive modeling based on chemical structure is already in use, known as quantitative structure–activity relationship modeling (QSAR). QSAR predicts irritation based on the known structure of the molecule. One such system is known as DEREK, which is discussed more in Chapter 28.

23.4  Clinical patterns of irritant contact dermatitis The clinical morphology of ICD is heterogeneous. Acutely, contact with irritants may produce erythematous and edematous patches or plaques with possible vesiculation in the location of exposure. If exposure to strong acids and alkalis has occurred, there may also be cutaneous ulceration due to the corrosive properties of these chemicals. The risk of ulceration increases with larger volumes of exposure, preceding trauma, and concurrent friction. With chronic exposure to less intense irritants, the patient may exhibit dry, fissured, and lichenified plaques. Less common morphologies of ICD include granulomas, folliculitis,

Hayes’ Handbook of Pesticide Toxicology

nonimmunologic urticaria (wheals), miliaria, and changes in pigmentation (Chew and Maibach, 2006). Chew and Maibach classified the more salient types of irritant contact dermatitis based on clinical presentation, chronology, and clinical course. A brief summary of these types follows.

23.4.1  Acute Irritant Dermatitis (Primary Irritation) This is the classic skin irritant response often seen as a result of exposure to a strong irritant such as a potent acid or alkaline solution. The skin responds immediately with erythema, edema, and possibly vesiculation, ulceration, and local necrosis (Slodownik et al., 2008; Welfriend and Maibach, 2008). Once the irritant is removed, the skin begins to heal. This is known as the decrescendo phenomenon. This is unlike allergic contact dermatitis, in which the inflammation increases after removal of the agent (crescendo phenomenon) before it eventually fades (Chew and Maibach, 2006).

23.4.2  Delayed, Acute Irritant Dermatitis This type of dermatitis is clinically similar to acute irritant dermatitis; however, it is characterized by a delayed onset of irritation after exposure. Generally, inflammation occurs 8–24 h after exposure and thus may mimic an allergic contact dermatitis (Chew and Maibach, 2006; Welfriend and Maibach, 2008). In these cases, thorough history, physical exam, and diagnostic patch testing can help distinguish between the two entities.

23.4.3  Irritant Reaction This type of reaction often arises in the first months of intense exposure to the irritant. Clinically, this condition is characterized by a monomorphous response. Individuals may display redness, scaling, vesicles, pustules, or erosions but not more than one characteristic. A classic example is occupations in which workers are exposed to wet work, such as beauty salon employees. Many of these workers have extensive water and soap exposures on a daily basis (i.e., during the shampooing process). Chronic exposures to the elements such as wind and cold can also result in an irritant reaction, such as dry lips on skiers. Often, this type of reaction heals without treatment (Chew and Maibach, 2006).

23.4.4  Subjective/Sensory Irritation In subjective/sensory irritation, there are symptoms of irritation such as stinging and burning without evidence of inflammation or damage clinically and histologically. The mechanism of this irritation is not known, although this

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must be differentiated from subclinical contact urticaria. One common chemical known to cause this phenomenon is lactic acid (Chew and Maibach, 2006; Welfriend and Maibach, 2008).

(Simon et al., 2001). This is also known as eczema craquele (Chew and Maibach, 2006).

23.4.5  Suberythematous Irritation

This dermatitis is characterized by erythema, scaling, and fissuring as a result of chronic low-grade friction. This should not be confused with skin thickening (i.e., callus formation or lichenification) that more commonly occurs due to chronic friction. Generally, these lesions improve with friction avoidance (Freeman, 2000).

This entity is characterized by early epidermal damage without visible inflammation. Clinically, the patient experiences burning, itching, and stinging (Chew and Maibach, 2006).

23.4.6  Cumulative Irritant Dermatitis This classification is likely one of the most prevalent morphologies seen by occupational physicians and dermatologists. Multiple exposures to weak irritants may eventually result in cutaneous irritation without an obvious acute dermatitis. The hallmark clinical features are erythema, dryness, scaliness, and eventual hyperkeratosis and skin fissures. Unlike acute irritant dermatitis, this dermatitis may arise over weeks to years. Patch testing helps separate this from its mimic, allergic contact dermatitis(Chew and Maibach, 2006).

23.4.7  Traumatic Irritant Dermatitis This peculiar type of irritant dermatitis occurs after acute trauma to the skin (Mathias, 1988). It may mimic cumulative irritant dermatitis or may present as nummular (coinshaped) erythematous patches and plaques (Welfriend and Maibach, 2008). A dyshidrotic eruption on the hands has also been described (Beukers and van der Valk, 2006). These lesions are often notably resistant to treatment, and they may take months to years to resolve (Mathias, 1988). The etiology is unknown, but in some cases it may be attributed to soaps and other topicals used to treat the wound (Slodownik et al., 2008).

23.4.8  Acneiform and Pustular Irritant Dermatitis Pustular irritant dermatitis often mimics the presentation of folliculitis or acne. Follicular-based erythematous papules and pustules occur in the area of irritant exposure. Classically, this response is seen after exposure to metals, oils, tar, asphalt, halogens, formaldehyde, aromatic hydrocarbons, chlorinated napththalene, and polyhalogenated naphthalene (Andersen and Petri, 1982; Chew and Maibach, 2006; Welfriend and Maibach, 2008).

23.4.9  Exsiccation Eczematoid Dry icthyosiform scaling, particularly in elderly individuals, is seen as a result of low humidity and cold temperatures

23.4.10  Friction Dermatitis

23.4.11  Airborne Irritant Dermatitis Airborne irritant dermatitis is similar to other types of acute and cumulative irritant dermatitis but has a characteristic clinical presentation. Because the irritant is in the air, the dermatitis generally involves uncovered skin, such as face, eyelids, arms, and V of the neck. This distribution is important to recognize in agricultural workers because aerosolized pesticides/fumigants may be cutaneous irritants (LaChapelle, 2006).

23.5  Diagnosis of irritant contact dermatitis The diagnosis of an acute corrosive-type irritant reaction is often self-evident. Generally, workers will recall exposure to a strong irritant chemical with immediate severe skin erythema, edema, vesiculation, or ulceration. Thus, the offending chemical is easily identified. Cases of cumulative irritant dermatitis are far more common in a clinical setting and more difficult to identify. These cases can be confused with allergic contact dermatitis, endogenous eczema, psoriasis, and other papulosquamous diseases. For these cases, thoughtful and complete history taking by the physician is crucial. One should inquire about the type of occupation and hobbies; daily activities within that occupation; and exposure to water, detergents, and other chemicals. Questions should be asked regarding exposure to physical irritants such as friction, low humidity, and heat. Attention to frequency and timing of these exposures is also important. Dermatitis that improves with time off from work suggests an occupational source. The distribution of lesions can be useful in identifying the source of the irritation. For example, an airborne irritant dermatitis such as a fumigant may present as a symmetric dermatitis on exposed areas of the body, especially eyelids and face (Dooms-Goossens et al., 1986), whereas harvesters may develop dermatitis on forearms and hands when in contact with irritant pesticide residues on foliage [Centers for Disease Control and Prevention (CDC), 1986]. Close physical inspection to exclude other common skin conditions, such as psoriasis, atopic dermatitis,

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and seborrheic dermatitis, is recommended. Table 23.1 presents the differential diagnosis of contact dermatitis and physical findings that are helpful in discriminating between these conditions. If an irritant or allergic contact dermatitis is suspected and one cannot distinguish between the two, patch testing is recommended. Patch testing is discussed in more detail in Chapter 25. Often, the diagnosis of ICD is one of exclusion – the clinical appearance suggestive of a contact dermatitis and negative patch tests for potentially relevant allergens favor a diagnosis of ICD.

23.6  Treatment of irritant contact dermatitis Identification and avoidance of the irritant is crucial in the treatment of irritant contact dermatitis. Use of protective devices such as gloves to prevent irritant exposure is an important preventive and treatment measure (Kwon et al., 2006). Corrosive reactions are best treated with irrigation (except in burning fragments of sodium, potassium, and lithium), specific antidotes for some chemicals, and local

wound/burn care. Having detailed knowledge of the product to which a person has been exposed can aid in the management of corrosive reactions (Bruze et al., 2006). For more cumulative ICD, emollients are often helpful when used frequently to improve skin barrier function. However, it appears that emollients may not be broadly effective in improving all patients with irritant dermatitis (Yokota and Maibach, 2006). Symptomatically, oral antihistamines may be helpful in preventing pruritus. For actively inflamed nonulcerated lesions, short courses of topical corticosteroids may be tried. Long-term treatment should be avoided to decrease the risk of skin atrophy and barrier dysfunction (Kao et al., 2003). The utility of topical corticosteroids for irritant dermatitis has been questioned. In one study of an acute experimental surfactant-induced irritant dermatitis, investigators found no improvement of lesions with use of low- and high-potency topical steroids (Levin et al., 2001). However, it is one author’s experience (L. M.) that short bursts of topical steroids are often worth an initial attempt to improve symptoms until avoidance strategies may be implemented. In addition, a few small studies suggest that topical calcineurin

Table 23.1  Differential Diagnosis of Contact Dermatitis Endogenous skin disease

Clinical features that may aid in diagnosis

References

Psoriasis

Well-demarcated scaly erythematous plaques on extensor surfaces, elbows, knees, scalp, and umbilicus Nail pitting Onycholysis and yellow nail discoloration Orange-yellow discoloration under nail (oil spots) Inflammatory arthritis and arthralgias

Griffiths and Barker (2007) Schon and Boehncke (2005)

Atopic dermatitis

Pruritic flexural erythematous papules, patches, and plaques Often located in popliteal fossae, antecubital fossa, and face Personal history of seasonal allergies or asthma Onset in early childhood Palmar hyperlinearity Icthyosis Nipple dermatitis Periorbital dermatitis with Dennie–Morgan line

Boguniewicz (2000)

Seborrheic dermatitis

Greasy or powdery scale in scalp, posterior auricular region, eyebrows, and nasolabial folds

Johnson and Nunley (2000)

Dermatophytosis

Scaly plaques that may be in an annular configuration Increased scale or pustules at leading edge Erythematous scaly plaques in moccasin distribution on feet Two plantar surfaces involved and one palm involved (two foot, one hand presentation) Yellowing and thickening of nails White crumbling nail surface Hyphal elements seen on potassium hydroxide preparation Culture positive for dermatophyte Hyphae seen with PAS stain on biopsy

Zuber and Baddam (2001)

Chapter | 23  Irritant Dermatitis

inhibitors may improve signs and symptoms of ICD (Engel et al., 2008; Mensing et al., 2008). For recalcitrant lesions, phototherapy or systemic immunosuppression with cyclosporine or azathioprine may be helpful (Cohen and Heidary, 2004).

23.7  ICD in agricultural workers 23.7.1  Acute Agricultural Irritant Dermatitis Agricultural work involves a complex and variable set of potential skin irritants. Depending on the season and climate and crop, these may include dust, heat, agricultural chemicals, and irritant chemicals derived from plants. For agricultural pest control workers, most of the irritant materials encountered may be synthetic pesticides and the hazard most obvious following accidental direct exposure. Between 1982 and 2006, the handler database included 1990 cases of possible, probable, and definite cases of skin reactions to single pesticide active ingredients in California pesticide handlers. (These included 653 skin reactions possibly related to pesticide application work, usually without direct exposure.) Although inert as well as active ingredients in pesticide formulations may cause skin irritation, for strongly irritant active ingredients the reported cases correlate with the results of experimental testing in animals and skin reactivity predicted from the DEREK model (see Chapter 28). SICRET (Skin Irritation Corrosion Rules Estimation Tool) is another model that helps predict the irritant potential of a chemical (Walker et al., 2005).

23.7.2  Acute Irritation from Pesticides (Fumigants and Insecticides) A large majority of the 149 reported definite and probable cases of irritant dermatitis associated with fumigants occurred in workers handling halogenated compounds and compounds releasing the irritant compound methyl isothiocyanate (MITC) (Table 23.2). There were 191 cases of probable and definite dermatitis associated with insecticide applications in California between 1982 and 2006. Data on two of the most frequently reported insecticides in this case series are given in Table 23.3. Propargite, despite institution of water-soluble bags during the 1970s, accounted for 47.6% of the cases. Its irritant properties are suggested by predictive modeling and also by results of animal tests. For the organophosphate chlorpyrifos, there were considerably fewer cases and less clear-cut results from Draize testing. It does not contain any of the reactive elements identified by the DEREK model. Additional data on the irritant properties of insecticides are discussed in Chapter 28.

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23.7.3  Cumulative Irritant Dermatitis from Pesticides As mentioned previously, because multiple exposures are required to manifest cumulative ICD, recognition of the offending agent may be challenging. Single case reports regarding potential pesticide-induced ICD have several potential explanations that require some effort to differentiate. Some cases might be reactions unique to a single crew member; some might be sentinel cases signifying the occurrence of an otherwise unrecognized outbreak; and some cases might prove to be nonwork-related skin conditions not expected, in most circumstances, to occur in coworkers. In investigating potential ICD cases and outbreaks, it is usually obvious whether the problem is work related. The central questions are whether the reported episode was related to pesticides and which material, among those reported, was principally responsible. A few examples of cumulative irritant dermatitis outbreaks have been described among fieldworkers in contact with pesticide residues on foliage. Repeated exposures to residues resulted in dermatitis. The persistence of pesticide residues depends on the amount of pesticide used, the halflife of the pesticide dissipation, the type of crop, and the type of work performed. Variation in residue dissipation is illustrated by data on propargite. Residue studies (Maddy et al., 1977, 1979) performed in a coastal area of California showed 1- or 2-day dissipation half-lives. Residue studies in California’s Central Valley typically showed half-lives of 5–7 days (Reeve et al., 1991), but some fields showed half-lives up to 11 days. Dissipation half-lives as long as 30 days have been measured in the context of outbreak investigations (O’Malley, 1998; O’Malley et al., 1989; Smith, 1991). To prevent exposure to an irritant residue, regulators must determine “safe-entry waiting periods” or the re-entry interval for the pesticide. The length of the required re-entry interval depends on both the irritant capacity of the individual compound but also on the level of initial residue deposition and the postapplication rate of residue dissipation. In 1988, an outbreak of dermatitis occurred among a crew of nectarine harvesters in Tulare County (Figure 23.1). Examination of a comparison group of workers who had dermatitis allowed an analysis of work history and residue history. This showed a strong correlation between cumulative exposure to propargite and the occurrence of dermatitis (Figure 23.2). A review of the work history for the group with dermatitis showed a peak exposure to propargite residues of 0.2 pg/cm2. This value was used as an estimated no-observed-effect residue level for purposes of determining a safe re-entry interval. The re-entry interval for harvesting tree fruit was lengthened to 21 days following the episode (O’Malley et al., 1990). In 1995, an outbreak of dermatitis on the chest, neck, arms, and face occurred among workers performing hand

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Table 23.2  Fumigant Irritants Probable definite cases in fumigant handlers, 1982–2006

No. of cases

Animal testing

Reactive structures identifiable based on DEREK and SICRET models

Halogenated fumigants

92

Methyl bromide

70

Reported as corrosive in public domain literature

1,3-Dichloropropene and D–D mixture

16

Multiple products (60.3% 1,3 dichloropropene and 3.2% chloropicrin; 81.2% 1,3-dichloropropene,16.5% chloropicrin) corrosive in the Draize assay; a 92% liquid formulation without chloropicrin caused minimal irritation

Ethylene dibromide

4

Reported as severe irritant in public domain literature

Methyl iodide

1

99.7% liquid technical material severe irritant in the Draize assay

MITC-releasing fumigants

51

Metam sodium

49

Five liquid products (32.7–43.8% metam sodium) corrosive in the Draize test; unexpected minimal irritation reported for three similar products (32.7–42.2% metam sodium)

Dazomet

2

24% liquid corrosive in Draize test; 2 20% liquid products and a 98.5% solid reported to cause minimal irritation

Other fumigants

7

Sulfuryl fluoride

5

Unable to perform Draize test because of physical properties of gas

Cases in applicators possibly related to rapid evaporation of liquid sulfuryl fluoride, akin to liquid nitrogen burns

Ethylene oxide

1

Unable to perform Draize test because of physical properties of gas

DEREK: IUNIQ ���������������������������� – Electrophile, generally no prolonged skin contact because of physical properties

DEREK IX: Reactive aliphatic halides, olefins SICRET: Halogenated alkanes and alkenes listed as potential skin irritants

DEREK: IUNIQ – Isocyanate strong nucleophile SICRET: Thiocyanates, cyanates not listed

SICRET: Epoxides considered potentially irritant or corrosive Aluminum phosphide

1

Unable to perform Draize test because of physical properties of phosphine gas

labor activities on a table grape ranch in northern Fresno County, California. Of 202 fieldworkers, 65 (32.2%) sought treatment for the dermatitis. The large number of workers involved suggested contact with an irritant rather than allergen. Several different pesticide residues were detected in these fields: sulfur, propargite, iprodione, myclobutanil, and dichloronitroaniline. Propargite was suspect because its direct irritant capacity was higher than that

Releases phosphine, weak base, electrophile relative to alkyl grignards (IUNIQ), but generally no prolonged skin contact because of physical properties

of other pesticides used on the field and it was found in higher than no-observed-effect levels. Some workers had a rapid onset of their dermatitis in relation to exposure, whereas others appeared to require a few days of cumulative exposure before the dermatitis developed. After further evaluation of propargite residue levels, it was presumed that slow dissipation of propargite was the cause of this outbreak (O’Malley, 1998).

Chapter | 23  Irritant Dermatitis

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Table 23.3  Insecticide Irritants Probable definite cases in insecticide handlers, 1982–2006

No. of cases

Animal testing

Reactive structures identifiable based on DEREK and SICRET models

Propargite

91

Technical material (listed as 90.6% AI) and the liquid formulation used on cotton (73.86% AI) caused corrosion in the Draize assay. The emulsifiable concentrate (69.62% AI) caused severe irritation. Two powdered formulations (28.99% AI and 32% AI) nevertheless were reported to cause minimal irritation in the Draize assay

DEREK: Terminal propargyl group containing unsaturated olefin (triple bond) SICRET: Alpha-alkynes likely to cause skin irritation

Chlorpyrifos

12

Technical chlorpyrifos (97.6% AI) caused transient irritation; some EC formulations with 40% AI caused moderate to severe irritation; dilute formulations with 1% AI all caused minimal irritation

No identified reactive moieties

Another example of cumulative dermatitis related to pesticides occurred in 1986, when a dermatitis outbreak occurred among orange pickers in California. Based on physician reporting, 58% of 198 workers developed a dermatitis involving most commonly the neck and the chest. Workers often leaned into foliage to pick the oranges, thus explaining the distribution of the dermatitis and suggesting pesticide residues as a possible cause. The miticide OMITE-CR was the suspected cause of the dermatitis because no cases of dermatitis among workers occurred prior to the application of OMITE-CR. Furthermore, there was a positive correlation between OMITE-CR residue hours (estimated leaf residue multiplied by hours of exposure) and the development of dermatitis (CDC, 1986). Figure 23.1  Variable onset of cumulative irritation in a crew of nectarine harvesters, June 1988 (reprinted with permission from Hanley and Belfus, State of the Art Reviews in Occupational Medicine, 1997).

Figure 23.2  Increasing cumulative incidence of dermatitis with progressive exposure to propargite (reprinted from O’Malley, 1997, with permission from Hanley and Belfus).

23.7.4  Plants as Agricultural Irritants Plants may also cause irritant dermatitis in the agricultural worker. Mechanistically, irritant contact dermatitis may arise from chemical and/or mechanical injury from the plant. Physical irritants include thorns, sharp leaves, spines, and irritant “hairs,” and these may produce a variety of dermatologic lesion morphologies. For example, contact with cactus spines may result in a pruritic papular eruption in the location of contact. However, some spines and thorns lodged in the skin may result in persistent foreign body granulomas that resemble other granulomatous diseases such as granuloma annulare. Moreover, spines may be a conduit for inoculation of infectious organisms into the skin (Lovell, 1993). Chemical injury may occur from a variety of plant compounds. One frequent offending irritant, calcium oxalate, is found in many plants, including agave, dumb cane, daffodils, and other “bulb flowers” such as hyacinth

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(Bruynzeel et al., 1993; Julian and Bowers, 1997; Lovell, 1993). Workers exposed to agave harvested to produce tequila may develop cutaneous lesions commonly on the forearms, neck, and abdomen, known as “mal de agaveros” (agave worker’s sickness) (Salinas et al., 2001). Similarly, those exposed to calcium oxalate and alkaloids in daffodils may develop an eczematous or granulomatous dermatitis

that affects the wrist and the fingers known as “daffodil pickers’ rash” (Julian and Bowers, 1997). Pineapples contain bromelin, a proteolytic enzyme, as well as calcium oxalate in their juice, both of which may cause irritant dermatitis (Bruynzeel et al., 1993; Fisher and Mitchell, 2001). Table 23.4 lists some other plant families that may cause irritant dermatitis. Phytophotodermatitis is a particular type

Table 23.4  Some Plants Known to Cause Irritant Dermatitis Plant

Characteristics

Contact reactions

Anthemis cotula (mayweed, dog fennel, camomile) Member of Compositae family

A species of weed introduced to the United States from Europe. It can be found sporadically throughout the United States. It is found on roadsides, orchards, pastures, and agricultural lands. Irritant found in the plant’s volatile oil

Rowe (1934) applied dry samples of the plant to the normal skin of 21 adults for 24 hours. Sixteen subjects showed definite areas of irritant dermatitis. Several workers pulling weeds manually in a field of winter sugar beets in California developed vesicular or blistering dermatitis due to contact with A. cotula (O’Malley et al., 2001). This was attributed to an irritant reaction May also cause allergic contact dermatitis (Menz and Winkelmann, 1987) and contact urticaria (Shelmire, 1940)

Cocklebur (Xanthium strumarium, Xanthium pennsylvanicum) Member of Compositae family

Common weed in the United States

Mechanical irritant; spines on fruit (Lovell, 1993). O’Malley et al. (2001) reported suspected irritant dermatitis in workers pulling weeds. The most prevalent weed was the cocklebur Cocklebur extract is known to cause irritant reactions (Mitchell et al., 1980) May also cause an allergic contact dermatitis (Menz and Winkelmann, 1987)

Velvet leaf (Abutilon theophrasti)

Common weed in the United States/Canada, particularly in the Midwest

O’Malley et al. (2001) reported cases of irritant dermatitis of hands and forearms in workers pulling weeds and encountering velvet leaf

Borage

Mass cultivated for oil

Physical irritant by penetration of skin by coarse “hairs” results in a papular irritant eruption (Lovell, 1993)

Euphorbia family (spurge)

Some in this family – E. pepulus (petty spurge), E. helioscopia (sun spurge), and E. lathyrus (caper spurge) – are weeds Contain irritant milky latex

Contact with latex may produce erythema and blistering on skin. May also cause irritant keratoconjunctivitis (Calnan, 1975; Lovell, 1993; Webster, 1986)

Ranunculaceae family (buttercup family)

The irritant protoanemonin is formed after injury to the plant. It is only found in freshly injured plants Can be found in field buttercups

May cause severe vesiculation mimicking a phototoxic reaction (Lovell, 1993; Oztas et al., 2006)

Brassicaciae family (radish, horseradish, and mustard)

Irritant is thiocyanate

Can cause irritation (Cleenwerke and Martin, 1995). May also cause allergic contact dermatitis (Mitchell and Jordan, 1974)

Peppers

Irritant is capsaicin

“Hunan hand” Workers who pick or otherwise handle hot peppers may be subject to burning, irritation, and erythema, without vesiculation (Williams et al., 1995)

Chapter | 23  Irritant Dermatitis

of chemical-induced irritant contact dermatitis that occurs after exposure to the offending plant and solar radiation. This entity is discussed in Chapter 24. Lastly, chemical toxins within the plant can be injected via physical means, as seen with members of the plant family Urticaceae (stinging nettle). These plants have small spines that contain histamine and produce wheals (urticaria) upon contact with the skin (Lovell, 1993). When evaluating a patient with suspected plant dermatitis, it should be noted that some plants may have the ability to cause irritant as well as allergic (immune-mediated) dermatitis.

Conclusion It would be ideal if agricultural workers knew the irritating potential of each chemical product encountered in their work. Unfortunately, most of the animal-based registration data remain unavailable to workers, and if available, few would be prepared to interpret the data. Few human studies exist, and new studies are currently inhibited by the U.S. Environmental Protection Agency’s ethics system. Lastly, epidemiologic data that are so helpful for many contact allergens remain scarce for agricultural irritants. The authors welcome any initiative that will help the worker. The techniques and assays are efficient; however, a regulatory system that promotes developing and registering relevant data in this arena is lacking.

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Rowe, A. (1934). Camomile (Anthemis cotula) as a skin irritant. J. Allergy 5(4), 383–388. Salinas, M. L. et al. (2001). Irritant contact dermatitis caused by needlelike calcium oxalate, crystals, raphides, in Agave tequilana among workers in tequila distilleries and agave plantations. Contact Dermatitis 44, 94–96. Santucci, B. et al. (2003). Cutaneous response to irritants. Contact Dermatitis 48, 69–73. Schon, M. P., and Boehncke, W. H. (2005). Psoriasis. N. Engl. J. Med. 352(18), 1899–1912. Shelmire, B. (1940). Contact dermatitis from vegetation. Patch testing and treatment with plant oleoresins. J. South. Med. Assoc. 33, 337–346. Simon, M. et al. (2001). Persistence of both peripheral and non-peripheral corneodesmosomes in the upper stratum corneum of winter xerosis skin versus only peripheral in normal skin. J. Invest. Dermatol. 116, 23–30. Slodownik, D. et al. (2008). Irritant contact dermatitis: a review. Australas. J. Dermatol. 49, 1–9 quiz 10-1. Smith, C. R. (1991). Dissipation of dislodgeable propargite residues on nectarine foliage. Bull. Environ. Contam. Toxicol. 46, 507–511. Thestrup-Pedersen, K. (2000). Clinical aspects of atopic dermatitis. Clin. Exp. Dermatol. 25, 535–543. Tupker, R. A. (2003). Prediction of irritancy in the human skin irritancy model and occupational setting. Contact Dermatitis 49, 61–69. Van der Valk, P. G., and Maibach, H. I. (1989). Post-application occlusion substantially increases the irritant response of the skin to repeated short-term sodium lauryl sulfate (SLS) exposure. Contact Dermatitis 21, 335–338.

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Walker, J. D. et al. (2005). The skin irritation corrosion rules estimation tool (SICRET). QSAR Combinatorial Sci. 24, 378–384. Watkins, S. A., and Maibach, H. I. (2009). The hardening phenomenon in irritant contact dermatitis: An interpretative update. Contact Dermatitis 60, 123–130. Webster, G. L. (1986). Plant dermatitis. Irritant plants in the spurge family (Euphorbiaceae).�� Clin. Dermatol. 4(2), 36–45. Welfriend, S., and Maibach, H. (2008). Irritant dermatitis: Clinical heterogeneity and contributing factors. In “Marzulli and Maibach’s Dermatotoxicology” (A. L. Chew and H. I. Maibach, eds.). CRC Press, Boca Raton, FL. Welss, T. et al. (2004). In vitro skin irritation: facts and future. State of the art review of mechanisms and models. Toxicol. In Vitro 18, 231–243. Wester, R. C., and Maibach, H. I. (1985). In vivo percutaneous absorption and decontamination of pesticides in humans. J. Toxicol. Environ. Health, 16, 25–37. Wigger-Alberti, W. et al. (1997). Predictive testing of metalworking fluids: a comparison of 2 cumulative human irritation models and correlation with epidemiological data. Contact Dermatitis 36, 14–20. Williams, S. R., Clark, R. F., and Dunford, J. V. (1995). Contact dermatitis associated with capsaicin: Hunan hand syndrome. Ann. Emerg. Med. 25(5), 713–715. Yokota, M., and Maibach, H. I. (2006). Moisturizer effect on irritant dermatitis: An overview. Contact Dermatitis 55, 65–72. Zuber, T. J., and Baddam, K. (2001). Superficial fungal infection of the skin. Postgrad. Med. 109(1), 117.

Chapter 24

Photocontact Dermatitis Mikael Langner1, Howard I. Maibach1 and Lisa E. Maier2 1 2

University of California, San Francisco, California University of Michigan, Ann Arbor, Michigan

24.1  Introduction Photosensitivity is an abnormal cutaneous reaction to solar ultraviolet radiation. This reaction may clinically manifest as greater propensity toward sunburn or development of a rash upon exposure to solar radiation. A variety of etiologies may be responsible, including porphyria, connective tissue disease, nutritional abnormalities, genetic diseases, and idiopathic disorders (Lim et al., 2007). Particularly relevant to the agricultural worker is photosensitivity as a result of contact with exogenous factors, such as plants, pesticides, and sunscreens. These exogenous photocontact reactions are known as photoallergy and phototoxicity (Marzulli et al., 2008). In this chapter, we focus on photo­allergy and phototoxicity and list some relevant compounds responsible for these distinct entities. We explain the clinical workup of photosensitive patients, including phototesting, and discuss management strategies for photosensitivity and photoallergy.

24.2  Solar radiation and photosensitivity Ultraviolet light is non-ionizing radiation on the electromagnetic spectrum between visible light and X-rays. Overall, ultraviolet radiation constitutes approximately 10% of the total solar radiation that reaches the earth, with infrared radiation and visible light constituting the majority of other solar radiation types (Moyal et al., 2004). UVA rays (wavelength, 320–400 nm) make up approximately 95% of the ultraviolet radiation (UVR) that reaches the earth, whereas UVB (wavelength, 290–320 nm) comprises the other 5% (Lim et al., 2007). UVA appears to be responsible for photosensitivity in most cases; however, exposure to UVB and even visible light may induce photosensitivity in the susceptible individual (Lim et al., 2007). Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

24.3  Photoirritant reactions (phototoxicity) Phototoxic reactions are the most common type of exo­ genous photosensitivity. This is a nonimmune-mediated mechanism of photosensitivity that may occur in any individual. In these reactions, an offending chemical absorbs UVR and becomes activated; this activated chemical then causes direct tissue damage. Without UVR, in most cases these chemicals are otherwise inert. Clinically, phototoxic reactions mimic robust sunburns in the areas exposed to the chemical and the UVR, and generally, phototoxic reactions occur quickly, over minutes to hours, whereas photoallergic reactions occur relatively slowly, over 24–72 h (Lugovic et al., 2007). Phototoxic reactions result from various compounds, including antibiotics (tetracyclines, fluoroquinolones, and sulfonamides), nonsteroidal anti-inflammatory drugs (NSAIDs), diuretics (furosemide and hydrochlorothiazide), cosmetics containing oleum bergamote, plants, and some pesticides (Lugovic et al., 2007). Phytophotodermatitis (PPD) is a specific type of photo­toxic reaction that results from exposure to UVA and photosensitizers, known as furocoumarins, found in plants (Derraick and Rademaker, 2007; Klaber, 2006). Furocoumarins include psoralens, angelicin, bergaptol, and xanthotal. Photosensitizer-containing plants have highest concentrations in the spring and summer when it is most common for patients to present with PPD (Derraick and Rademaker, 2007). PPD has a distinctive clinical presentation (Figure 24.1), characterized by edematous and erythematous patches that frequently develop vesicles or bullae. These patches are often in linear or angular configurations on areas where the worker has brushed against the plant. The dermatosis subsides within 1 or 2 weeks but frequently results in persistent hyperpigmentation (Juckett, 1996). Other configurations of the dermatitis may occur, depending on the location of contact. For example, children 661

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Table 24.1  Plant Families and Representative Plants Causing PPD

Figure 24.1  Patient with phytophotodermatitis after exposure to fig tree. Note band of sparing from watchband (reproduced with permission from Derraick and Rademaker, 2007).

making pea shooters from giant hogweed developed perioral blistering (Lovell, 1993), and bartenders may develop PPD on their fingers from squeezing limes (Pathak, 1986). PPD from systemic ingestion of natural furocoumarins is uncommon and appears to require consumption of large quantities of the photosensitizing plant. For example, a healthy 65-year-old vegetarian woman in Sweden consumed copious quantities of celery as part of her normal diet 1 h prior to a tanning salon visit. Subsequent to the tanning session, however, the woman developed a severe, generalized phototoxic reaction comparable to having received photochemotherapeutic doses of methoxsalen and UVA light (Ljunggren, 1990). Several plant families can cause phytophotodermatitis, and Table 24.1 provides a representative list. Although parthenium is capable of producing PPD due to thiophenes and acetylenes, it is most notorious for causing an airborne ACD (Jovanovic et al., 2004). During World War II, the Pune province of India was accidentally sent parthenium-infested grain to address food shortages. Since then, thousands of people have developed dermatitis due to parthenium exposure (Agarwal and D’Souza, 2009; Shaikh and Shaikh, 2008). Indeed, chronic sunlight exposure coupled with chronic photosensitizer exposure can lead to chronic actinic dermatitis, a chronic skin disorder fairly common in people older than 60 years (Lim et al., 2007). For these reasons, parthenium is of historical interest, particularly in Pune. Skin exposed to discoloring plant dyes can result in persistent hyperpigmentation; when considering PPD as a diagnosis, it is important to rule out this possibility by taking

Plant family

Specific plants

Umbelliferae (Apiaceae)

Celery Carrot Fennel Parsnip Giant hogweed Cow parsley

Rutaceae

Rue Burning bush Skimmias Citrus: orange, lemon, lime, grapefruit, bergamot orange

Moraceae

Fig tree (Ficus carica)

Asteraceae

Parthenium

Fabaceae

Babchi Scurf pea

From Derraick (2007).

Table 24.2  Botanical Agents Causing Hyperpigmentation Botanical

Type of pigmentation

Lime, lemon, orange juniper

Red discoloration by terpenes

Cinnamic alcohol

Postinflammatory hyperpigmentation or depigmentation

Geraniol, cananga oil

Postinflammatory hyperpigmentation

the patient’s history and asking about exposure to the offending botanical agents (White, 1996). Table 24.2 lists botanical agents that can cause hyperpigmentation; interestingly, cinnamic alcohol is capable of producing either depigmentation or hyperpigmentation.

24.4  Photoallergic contact dermatitis Photoallergic reactions result when a skin allergen is activated by light. This process involves a cell-mediated hypersensitivity reaction to the allergen, which, when activated by light, produces an immediate hypersensitivity reaction with rising IgE levels (Lim et al., 2007). The putative allergen may be a topically applied or systemically

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Table 24.3  Compounds Causing Photoallergic Response References

Study type

Occupation

Compound causing photoallergy

Darvay et al. (2001)

Retrospective (N  2715)

Various

Sunblock: benzophenone-3, benzophenone-10

Zeeli et al. (2006)

Photo-patch test series

Various

Fragrances

Ljunggren (1977)

Case report

Sales manager

Light  hogweed

Yazici et al. (2004)

Case report

?

Celecoxib

Karimian-Teherani et al. (2008)

Case report

Gardener

Heracleum giganteum

Serrano et al. (2008)

Case series

Woodmaker, carpenter

Brosimum wood

Ljunggren (1977)

Case report

Gardener

Parsley

Autio et al. (2004), Koch and Bahmer (1989)

Case report

Agricultural worker

Phenmedipham (a herbicide used in sugar beet cultivation)

Derraik (2007)

Case series of 20 patients

Various

Toxicodendron succedaneum

absorbed chemical or its metabolite; Table 24.3 gives several examples of photoallergens. Such uncommon reactions affect less than 1% of the population. It is noteworthy, however, that more than 50% of patients suffering from photoallergic reactions may also have an underlying photodermatosis, either polymorphic light eruption (PMLE) or chronic actinic dermatitis (Darvay et al., 2001; Zeeli et al., 2006). Epidemiologic information indicates that PMLE occurs more commonly in individuals younger than 30 years, whereas chronic actinic dermatitis has its onset more often in people older than 60 years (Lim et al., 2007). The clinical manifestation of photoallergic contact dermatitis is an eczematous type of cutaneous reaction. Erythematous patches and vesicles may be present in the acute stages; with repeated contact with the allergen, however, the patient may develop dry scaly patches and plaques with possible lichenification. Uncommonly, photoallergic contact dermatitis can present as a lichenoid dermatitis, characterized by erythematous to violaceous discrete papules and plaques in a sun-exposed distribution (Parodi et al., 1987; Verma et al., 2002). Pruritus is often prominent. These reactions are delayed, and they may occur 24–72 h after exposure to UVR (Lugovic et al., 2007). Figures 24.2 and 24.3 are representative of photoallergic contact dermatitis. Some of the most clinically relevant photoallergens are sunblocks; however, they have low rates of sensitization and photosensitization (Darvay et al., 2001). Two types of sunblock are available: the physical UV filters and the chemical UV filters. The physical UV filters – zinc and titanium oxide – do not induce photoallergy, but the chemical UV filters – benzophenone-3 and benzophenone-10 – can induce photoallergy (Darvay et al., 2001). Plants, topically applied fragrances, and topical systemic medications

Figure 24.2  Patient with photoallergic contact dermatitis after exposure to Heracleum giganteum (giant bear claw) (reproduced with permsission from Karimian-Teherani et al., 2008).

may also induce photoallergy. Several classes of drugs that are commonly associated with photoallergy include thiazide diuretics, sulfonamide antibiotics, sulfonylureas, and phenothiazines, all of which contain a sulfur moiety. Additional medications that have been reported as photoallergens include quinine, quinidine, tricyclic antidepressants, antimalarials, and NSAIDs. The list of drugs with photoallergic potential is increasing in rough proportion

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Figure 24.3  Photoallergic contact dermatitis after exposure to brosimum wood (reproduced with permission from Serrano et al., 2008).

to the number of new drugs each year (Leung and Greaves, 2000). Although there is overlap between compounds causing phototoxicity and photoallergy, much less compound is required for photoallergic reactions because the immune system presumably amplifies the effect of the compound on the body (Lugovic et al., 2007). In general, photoallergic reactions resolve if the diagnosis is made promptly and the offending agent is discontinued. Chronic exposure to photoallergic drugs can occasionally lead to extreme photosensitivity that persists for months to years even after the responsible drug is eliminated (Bolognia et al., 2007). This condition is known as persistent light reaction, and it must be included in the differential diagnosis of photosensitivity disorders. Finally, because exposure to even a tiny amount of the compound can trigger a photoallergic reaction, strict avoidance is often required (Lugovic et al., 2007).

24.5  Pesticides that cause phototoxic or photoallergic reactions The potential of pesticides to cause photocontact reactions is of importance to the agricultural worker. Pesticides with the ability to cause phototoxic or photoallergic reactions include the fungicide chlorothalonil (Daconil, Bravo, Echo, Exotherm Termil, Forturf, Mold-Ex, and Nopcocide N-96), a commonly used pesticide that can produce phototoxic skin eruptions upon light exposure and skin contact (Matsushita et al., 1996). The pesticides folpet and captan have been reported to induce photoallergic reactions (Mark et al., 1999). We suspect that photoirritation and/or photoallergic contact dermatitis may be seen with some other common pesticides, such as bithionol, dichlorophene,

Hayes’ Handbook of Pesticide Toxicology

dimethylol dimethyl hydantoin, and fenticlor (Lim, 2009). Sulfur and atrazine, two of the most commonly used pesticides in the United States, do not produce phototoxicity and photoallergic reactions. Until registration authorities mandate dermatotoxicologic assays of suspect agricultural chemicals, it will remain difficult to not only know the potential of these chemicals to cause contact or photocontact reactions but also to issue suitable health and safety precautions. Furthermore, until we deliver more complete diagnostic testing methods in pesticide-exposed workers, we will not know what has been missed. Although data from animal models are probably acceptable to extrapolate to human cases when the allergic reaction is severe, data are unavailable for many chemical structures because experiments are lacking, incomplete, or inconclusive (Penagos et al., 2000). Indeed, the history of dermatology is replete with missed diagnoses because of lack of dermatotoxicity assays and incomplete patient and worker diagnostic workup.

24.6  Approach to the photosensitive patient One should suspect a photosensitive disorder when there is cutaneous sensitivity in photoexposed areas such as the forehead, bridge of the nose, chin, malar cheek, the V area of the neck, and extensor extremities. Sometimes the relationship with sunlight is not completely obvious to the patient because responses can be delayed or occur after cumulative sun exposure, so a high index of suspicion is needed (Lim et al., 2007). Differentiating photocontact and phototoxic disorders from other photosensitive disorders may often be achieved by history, physical examination, laboratory evaluation, phototesting, and skin biopsy for histopathology (Yashar and Lim, 2003). Table 24.4 summarizes the appropriate workup. Often, there are specific clinical features that may steer the clinician toward a specific diagnosis. For example, PMLE is characterized by papules, vesicles, or plaques that develop generally with the first sun exposure, often in the spring or early summer (Bolognia et al., 2007), and most commonly in patients younger than 30 years. Salient features of other photosensitive disorders are listed in Table 24.5. Airborne contact dermatitis deserves special mention. This condition may mimic a photosensitive dermatosis, particularly photoallergy, with patients having a diffuse eczematous reaction in areas not covered by clothing (i.e., face, chest, neck, and arms – the same locations areas that are also exposed to solar radiation). Occasionally, a history of airborne allergen can be elicited; however, photo-patch testing may be required to distinguish from photocontact dermatosis. Exposure to the insecticide phoxim should raise clinical suspicion for airborne contact dermatitis (Nakamura and Miyachi, 2003). Furthermore, some

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Table 24.4  Clinical Workup of Photosensitive Patient Workup component Specific question or test History

Age of onset, typical eruption (e.g., time interval from exposure, duration, and season), family/personal history to include connective tissue or autoimmune disease, and occupation

Physical exam

Distribution of lesions in sun-exposed areas

Phototests

UVA, UVB, visible light, patch, and photo-patch testing

Labs

ANA panel, porphyrin profile, anti-SSB (La) and anti-SSA (Ro) titers, patch testing to appropriately diluted allergens, photo-patch testing

Histopathology

Skin biopsy (Yap et al., 2003); often nonspecific

Table 24.5  Photosensitive Disorders, Clinical Features, and Therapy Photosensitive disorder

Clinical features

Therapy

Polymorphous light eruption

Papules, vesicles, or plaques within hours of light exposure; resolution with desensitization treatment with PUVA and intermittent sun exposure is retrospectively pathognomonic (Lebwohl, 2005)

Sun protection, corticoids, narrowband UVB, PUVA, antimalarials

Systemic lupus erythematosus, a photoaggravated disorder (Bolognia et al., 2007)

Malar rash, discoid rash, photosensitivity, oral ulcers, arthritis, renal disorder, neurological disorder, hematologic disorder, immunologic disorder, anti-nuclear antibody

Sun protection, topical corticoids, calcineurin inhibitors, retinoids, systemic corticoids

Phototoxic contact dermatitis

Exaggerated sunburn-like response, postinflammatory hyperpigmentation. If phytophotodermatitis, may have linear or angular configuration

Sun protection, avoidance of phototoxic agent (Bolognia et al., 2007)

Actinic prurigo

Pruritic, papular, or nodular eruption on uncovered skin exposed to light

Sun protection, topical corticoids, calcineurin inhibitors (Bolognia et al., 2007)

Photoallergic contact dermatitis

Eczema-like eruption, lichen planus

Sun protection, avoidance of allergic trigger (Bolognia et al., 2007)

Porphyria cutanea tarda

Increased skin fragility, blistering, erosions, crusts, milia, and scars in sun-exposed areas; uroporphyrinogn decarboxylase mutation (Lebwohl, 2005)

Sun protection, avoidance of triggering factors such as alcohol, phlebotomy

Erythropoietic protoporphyria

Intense burning, stinging, and pruritus; ferrochelatase mutation (Lebwohl, 2005)

Sun protection, beta-carotene reduces free radical levels

Airborne contact dermatitis

Dermatitis in areas not protected by clothing; often occupational exposure (patient history key to differentiate from photosensitivity)

Avoidance of trigger (e.g., exposure to the insecticide phoxim) (Nakamura and Miyachi, 2003)

Chronic actinic dermatitis (aka persistent light reaction)

Erythematous dermatitis due to very short light exposure years after the photosensitizing chemical exposure

Avoid light exposure – even 2 or 3 min can trigger reaction; avoidance of relevant contact allergens (Bolognia et al., 2007)

From Bolognia et al. (2007).

patients with an airborne contact dermatitis to sesquiterpene lactones, found in the Compositae (Asteraceae) group of plants, will develop photosensitivity (Dooms-Goossens and Deleu, 1991; Moseley et al., 2009). If persistent, this dermatitis and photosensitivity may be termed chronic actinic dermatitis, as shown in Figure 24.4.

24.7  Phototesting Phototesting assesses cutaneous response to UVA, UVB, and visible light. On exposure to UVA or UVB light, the presence of erythema, vesiculation, bullae, or hyperpigmentation represents a positive test; on exposure to visible

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Table 24.6  Skin Types with Propensity to Burn and/or Tan in First 30–45 Minutes of Sun Exposure after a Winter Season of No Sun Exposure

Figure 24.4  Chronic actinic dermatitis (reproduced with permission from Yashar and Lim, 2003).

light, the presence of urticaria represents a positive test. The differential diagnosis of photosensitivity is wide, and appropriate workup depends on clinical suspicion for specific dermatitides. For example, if contact dermatitis is suspected, then patch testing to appropriately diluted antigens is appropriate (Lebwohl, 2005). Phototesting for photosensitivity involves either the phototest or the photo-patch test (Marzulli et al., 2008). As previously described, phototesting assesses cutaneous response to UVA, UVB, and visible light upon exposing a small area of the patient’s skin to an artificial light source. In this method, a minimal erythema dose (MED) for the patient is initially determined. The MED is defined as the lowest dose of irradiation that will cause minimally perceptible redness 24 hours after exposure to either UVA or UVB irradiation (Bolognia et al., 2007). The MED will differ according to the patient’s unique skin pigmentation profile. Table 24.6 describes the skin types and their propensity to tan or burn. If an endogenous photosensitive disorder or ingested photosensitizer substance is present, MED will be decreased, and edema, vesiculation, bullae, or hyperpigmentation on exposure to the light may occur. If photoallergy is suspected, then photo-patch testing is appropriate. A photo-patch test employs an ultraviolet radiation source and patch test materials. The chemicals suspected of causing the reaction are placed on patches, and then two sets of patches are applied to the patient’s back for 48 hours. After removal, one set is irradiated with UVA at a dose below MED and the other is protected from UV dose. The results are read at 48 and 96 hours. A positive reaction is defined by erythema, edema, or vesiculation. If there is reaction only at the irradiated site, then photoallergy is suspected. If there is reaction at both sites, then one suspects contact allergy. If there is reaction at both sites, but a stronger reaction at the irradiated site, then one suspects both contact allergy and photoallergy (Marzulli et al., 2008).

Skin type

Burns

Tans

I (sensitive)

Always

Never

II (sensitive)

Always

Minimally

III (normal)

Moderately

Gradually (light brown)

IV (normal)

Minimally

Well (moderate brown)

V (insensitive)

Rarely

Profusely (dark brown)

VI (insensitive)

Never

Deeply pigmented

From the Food and Drug Administration (1999).

24.8  Histology Another aid in the evaluation of the photosensitive patient is the histopathology. It may be used to distinguish among photoallergy, phototoxicity, and other photosensitive disorders, such as porphyria and chronic actinic dermatitis. For example, photoallergy often demonstrates spongiosis, an inflammatory intercellular edema of the epidermis. Phototoxicity may demonstrate spongiosis and epidermal keratinocyte necrosis (Bansal et al., 2006; Bolognia et al., 2007), whereas spongiosis and atypical mononuclear cells in the dermis and epidermis are suggestive of the reticuloid variant of chronic actinic dermatitis (Yashar and Lim, 2003). Finally, cutaneous porphyrias have dermatopathology features such as cell-poor subepidermal blisters and immunoglobulins and complement at the dermoepidermal junction and perivascular areas (Yashar and Lim, 2003).

24.9  Management strategies Identification and avoidance of further contact with the offending allergen or toxin and any potentially cross-reacting agents is paramount in the management of photosensitive individuals. Topical steroids and systemic antihistamines are the mainstays of treatment (Bolognia et al., 2007). Patch testing is the first step to elucidating potential triggers. Reduction to UV light exposure is the first-line therapy for photoallergy and chronic actinic dermatitis, although support for this recommendation comes from clinical trials with fewer than 20 subjects (Lebwohl, 2005). A trial of cyclosporine is second-line therapy supported by clinical trial data with more than 20 subjects. Finally, other therapies, including tacrolimus (a topical calcineurin inhibitor), sunscreens, azathioprine, prednisolone, and PUVA/UVB, are reportedly useful, but their use is limited by sparse clinical data.

Chapter | 24  Photocontact Dermatitis

We suspect that photoirritant and photoallergic formulations exist. It is hoped that inclusion of this chapter in this standard text will motivate industry regulators to require premarketing screening and postmarketing dermato­ allergy epidemiology. Methods exist for epidemiologic studies (Penagos et al., 2000), and with these data, dermatologists will be better able to identify and risk-manage such formulations.

Conclusion Photocontact dermatitis results from skin exposure to an endogenous or exogenous offending substance, and skin exposure to UV radiation worsens the condition. Phototoxic reactions are the most common type of exogenous reaction. PPD is a specific type of phototoxic exogenous reaction resulting from skin exposure to furocoumarins (commonly found in plants) and UVA radiation. Photoallergic reactions result when a skin allergen is activated by light. Although they are generally rare compared to other photocontact dermatoses, sunscreens made from chemical UV filters can uncommonly produce photoallergic reactions. Pesticides such as chlorothalonil can cause phototoxic or photoallergic reactions. Until dermatotoxicologic assays are done for many of the suspect agricultural chemicals, however, the field of dermatology will misdiagnose or completely miss the diagnosis of many photocontact dermatoses. Therefore, dermatotoxicologic characterization and identification is of the essence to susceptible individuals because photocontact dermatoses often resolve upon avoidance of the offending agent.

References Agarwal, K. K., and D’Souza, M. (2009). Airborne contact dermatitis induced by parthenium: A study of 50 cases in South India. Clin. Exp. Dermatol. 34, e4–e6. Autio, S., Siimes, K., Laitinen, P., Ramo, S., Oinonen, S., and Eronen, L. (2004). Adsorption of sugar beet herbicides to Finnish soils. Chemosphere 55, 215–226. Bansal, I., Kerr, H., Janiga, J. J., Qureshi, H. S., Chaffins, M., Lim, H. W., and Ormsby, A. (2006). Pinpoint papular variant of polymorphous light eruption: clinical and pathological correlation. J. Eur. Acad. Dermatol. Venereol. 20, 406–410. Bolognia, J., Jorizzo, J. L., and Rapini, R. P. (2007). “Dermatology.” Mosby, St. Louis. Darvay, A., White, I. R., Rycroft, R. J., Jones, A. B., Hawk, J. L., and McFadden, J. P. (2001). Photoallergic contact dermatitis is uncommon. Br. J. Dermatol. 145, 597–601. Derraik, J. (2007). Heracleum mantegazzianum and Toxicodendron succedaneum: Plants of human health significance in New Zealand and the National Pest Plant Accord. N. Z. Med. J. 120, U2657. Derraick, J., and Rademaker, M. (2007). Phytophotodermatitis caused by contact with a fig tree (Ficus carica). N. Z. Med. J. 120(1259).

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Dooms-Goossens, A., and Deleu, H. (1991). Airborne contact dermatitis: an update. Contact Dermatitis 25, 211–217. Food and Drug Administration (1999). Sunscreen drug products for over-the-counter human use; Final monograph. Fed. Reg. 64(98), 27666–27693. Jovanovic, M., Poljacki, M., Duran, V., Vujanovic, L., Sente, R., and Stojanovic, S. (2004). Contact allergy to Compositae plants in patients with atopic dermatitis. Med. Pregl. 57, 209–218. Juckett, G. (1996). Plant dermatitis. Possible culprits go far beyond poison ivy. Postgrad. Med. 100, 159–163, 167–171. Karimian-Teherani, D., Kinaciyan, T., and Tanew, A. (2008). Photoallergic contact dermatitis to Heracleum giganteum. Photodermatol. Photoimmunol. Photomed. 24, 99–101. Klaber, R. (2006). Phytophotodermatitis. Arch. Dis. Child. 91, 385. Koch, P., and Bahmer, F. A. (1989). Photoallergic dermatitis caused by the herbicide phenmedipham. Derm. Beruf Umwelt. 37, 203–205. Lebwohl, M. (2005). “Treatment of Skin Disease: Comprehensive Therapeutic Strategies.” Mosby, St. Louis. Leung, D., and Greaves, G. M. (2000). “Allergic Skin Disease.” Dekker, New York. Lim, H., Honigsmann, H., and Hawk, J. (2007). “Photodermatology.” Informa Healthcare, New York. Lim, H. W. (2009). Abnormal responses to ultraviolet radiation: Photosensitivity induced by exogenous agents. In “Fitzpatrick’s Dermatology in General Medicine” (G. L. Wolff K, S. I. Katz, B. Gilchrest, A. S. Paller, and D. J. Leffell, eds.). McGraw-Hill, New York. Ljunggren, B. (1977). Psoralen photoallergy caused by plant contact. Contact Dermatitis 3, 85–90. Ljunggren, B. (1990). Severe phototoxic burn following celery ingestion. Arch. Dermatol. 126, 1334–1336. Lovell, C. (1993). “Plants and the Skin.” Blackwell, Boston. Lugovic, L., Situm, M., Ozanic-Bulic, S., and Sjerobabski-Masnec, I. (2007). Phototoxic and photoallergic skin reactions. Coll. Antropol. 31(suppl 1), 63–67. Mark, K. A., Brancaccio, R. R., Soter, N. A., and Cohen, D. E. (1999). Allergic contact and photoallergic contact dermatitis to plant and pesticide allergens. Arch. Dermatol. 135, 67–70. Marzulli, F., Zhai, H., Maibach, H. I., Klaus, P. W., and Wilhelm, K. (2008). “Marzulli and Maibach’s Dermatotoxicology.” CRC Press, Boca Raton, FL. Matsushita, S., Kanekura, T., Saruwatari, K., and Kanzaki, T. (1996). Photoallergic contact dermatitis due to Daconil. Contact Dermatitis 35, 115–116. Moseley, H., Naasan, H., Dawe, R. S., Woods, J., and Ferguson, J. (2009). Population reference intervals for minimal erythemal doses in monochromator phototesting. Photodermatol. Photoimmunol. Photomed. 25, 8–11. Moyal D, F. A. (2004). Acute and chronic effects of UV on skin. What are they and how to study them? In “Photoaging” (W. R. Rigel DS, H. W. Lim, and J. S. Dover, eds.), Vol. 1. Dekker, New York. Nakamura, M., and Miyachi, Y. (2003). Airborne photocontact dermatitis due to the insecticide phoxim. Contact Dermatitis 49, 105–106. Parodi, G., Guarrera, M., and Rebora, A. (1987). Lichenoid photocontact dermatitis to musk ambrette. Contact Dermatitis 16, 136–138. Pathak, M. (1986). Phytophotodermatitis. Clin. Dermatol. 4, 102–121. Penagos, H., O’Malley, M., and Maibach, H. I. (2000). “Pesticide Dermatoses.” Informa Healthcare, Boca Raton, FL. Serrano, P., Medeiros, S., Quilho, T., Santos, R., and Brandao, F. M. (2008). Photoallergic contact dermatitis to brosimum wood. Contact Dermatitis 58, 243–245.

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Shaikh, W. A., and Shaikh, S. W. (2008). Allergies in India: An analysis of 3389 patients attending an allergy clinic in Mumbai, India. J. Indian Med. Assoc. 106, 220 222–224. Verma, K. K., Sirka, C. S., Ramam, M., and Sharma, V. K. (2002). Parthenium dermatitis presenting as photosensitive lichenoid eruption. A new clinical variant. Contact Dermatitis 46, 286–289. White, I. (1996). Plant products in perfumes and cosmetics. Semin. Dermatol. 15, 78–82. Yap, L., Foley, P., Crouch, R., and Baker, C. (2003). Chronic actinic dermatitis: a retrospective analysis of 44 cases referred to an Australian photobiology clinic. Australas. J. Dermatol. 44, 256–262.

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Yashar, S., and Lim, H. W. (2003). Classification and evaluation of photodermatoses. Dermatol. Ther. 16, 1–7. Yazici, A., Baz, K., Ikizoglu, G., Kokturk, A., Uzumlu, H., and Tataroglu, C. (2004). Celecoxib-induced photoallergic drug eruption. Int. J. Dermatol. 43, 459–461. Zeeli, T., David, M., and Trattner, A. (2006). Photopatch tests: any news under the sun? Contact Dermatitis 55, 305–307.

Chapter 25

Allergic Contact Dermatitis

Sara Flores and Howard Maibach University of California, San Francisco, California

25.1  Introduction National statistics have consistently identified agriculture as an industrial division at the highest risk of occupational skin disease (O’Malley and Mathias, 1988). Allergic contact dermatitis (ACD) is one of the immune responses to agricultural chemicals demonstrated in the skin. Attempts have been made to access prevalence in the general population, but few data exist concerning the prevalence of ACD in any agricultural populations, including predominantly Hispanic California farm workers (Gamsky et al., 1992). In the past 25 years, much has been learned about mechanisms of immune responses in the skin, and our ability to predict agrochemical ACD has improved with introduction of predictive models. Cases and epidemiologic reports also contribute insight into environmental and physical conditions contributing to sensitization. This chapter provides a brief introduction to the advances and understanding associated with ACD and the relationship of cases to agrochemicals.

25.2  Allergic contact dermatitis Contact dermatitis is an important cause of occupational skin diseases and accounts for 15–20% of all cases reported (Smith and Hotchkiss, 2001). The skin reactions belonging to contact dermatitis include irritant dermatitis and allergic contact dermatitis. Irritant dermatitis (ID) results from activation of innate immunity and represents a nonspecific response from the immune system to a chemical or mechanical injury. It does not therefore involve antibody receptor specificity and usually corresponds to a dysfunction in the skin barriers. For example, frequent hand washing predisposes the skin Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

to irritant reactions because it damages the epidermal barriers (Jacob and Steele, 2006). The differences between irritant and allergic contact dermatitis are viewed in terms of concentration and frequency of exposure as well as the immune mechanisms involved. An acute irritant is a chemical that is presented in high enough concentration to elicit an immediate redness and inflammation at the site of exposure. In the initial stages following exposure to a small molecule, or xenobiotic, ID appears to be a distinct process from those of hypersensitivity reactions in that different cytokines are expressed (Muller et al., 1996). Allergic contact dermatitis is a skin reaction characterized by erythema, papules, and vesicles resulting from delayed-type hypersensitivity (DTH). As a type IV hypersensitivity reaction, ACD differs from ID in that it is mediated by T cells and characterized by the development of immunologic memory to the allergen. Another difference associated with the pathology for ACD is that it requires previous sensitization and is relatively antigen specific. Sensitization usually results from repeated use or exposure to an allergen, which is a chemical in smaller concentrations capable of triggering the characteristic immune response. The inflammation manifested is called contact sensitivity (CS) or contact sensitization and can result from contact with numerous agents. Many of these agents are low-molecular-weight chemical compounds. These compounds may be innately immuno­ genic, but they require some alterations before directing a response. The main one is the binding of the hapten to an epidermal keratinocyte, which marks the beginning of a multistep process that defines DTH reactions and ACD. Haptens are small enough to be absorbed into the skin, where they pass through the stratum corneum and bind with keratinocytes in the stratum spinosum below. This binding of

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hapten to keratinocytes forms a structure that is then a marker recognized by the immune system. Langerhans cells are dendritic cells specialized for presenting antigens by means of an MHC class II receptor to other immunologic cells. The Langerhans cells detect and phagocytize the unusual structure and then load the hapten onto an MHC class II receptor. The receptor travels back to the surface of the cell, where it displays the antigen as a hapten–carrier complex. The migration of the Langerhans cells to local lymph nodes exposes helper T cells to the antigen and triggers clonal expansion. The exact mechanism for this migration is not clearly understood, but it appears that inflammation caused by hapten invasion of the skin causes release of cytokines, which initiates their maturation and migration (Saint-Mezard et al., 2004). The activated T cells then travel in the circulatory and lymph systems to the skin, where they remain until subsequent exposure. This ends the first part of the immune response in ACD, which is called the sensitization phase. This creation of antigenspecific memory T cells occurs in 10–14 days. Studies have begun to differentiate between certain aspects of CS and DTH in that CS appears to involve the activation of both CD8 and CD4 T cells by Langerhans and other dendritic cells within the skin. In other words, CS appears to be more complex than a DTH reaction and can be started by either MHC class I presentation to CD8 T cells or by MHC class II presentation to CD4 T cells (Krasteva et al., 1999). In addition, the CD8 and CD4 T cells may have alternate roles in afferent and efferent parts of the second phase, the elicitation phase. The elicitation phase occurs upon subsequent exposure to the allergen. Symptoms are visible within 24–72 h because T cells recognizing the antigen are already present in the dermis and epidermis and are triggered more quickly by the hapten–protein carrier complex presented on the surfaces of Langerhans cells. The T cell receptor binds to the MHC receptor containing the hapten and causes the release of cytokines interleukin-1 (IL-1), IL-2, interferon, and tumor necrosis factor-, which are partially responsible for the inflammatory response and cutaneous lesions. The elicitation phase is further divided into two processes, the afferent process or phase and the efferent process/ phase. CD8 T cells are responsible for the afferent phase of the challenge reaction, whereas CD4 T cells are responsible for the efferent phase. The afferent phase is characterized by the release of cytokines mentioned previously. The cytokines initiate production of adhesion molecules within the capillaries. These adhesion molecules act to slow the movement of leukocytes and aid their transfer from the blood vessel to the dermis. They can then travel to the epidermis and induce the characteristic symptoms of ACD, mainly edema. The efferent phase is still not completely understood, and scientists are unsure how ACD spontaneously subsides. It usually lasts for approximately 72 h, and inflammation can persist for a few days.

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25.3  Prevalence of allergic contact dermatitis due to agrochemicals Assessing prevalence of ACD in the general population is not an easy undertaking. Diagnosis and differentiation between ACD and ID is difficult, and ACD heals spontaneously in days to weeks without medical intervention. Thus, many choose not to seek the advice of a professional, and there may be misdiagnoses with those who do. In addition, pesticide surveillance reports come from numerous sources. Many do not include information on ACD and are oriented either to systemic poisonings or to the tendency toward grouping all skin reactions into a more general classification, such as contact dermatitis. Furthermore, the prevalence of contact allergy differs among various countries, and interpretation and tests used in different locations or by different health care workers may contrast. Results are also affected by climate, culture, and cultural habits. For example, sensitization from direct contact during agrochemical applications may not be common among Americans because the majority do not handle them directly. Add to these factors the distinction between cases involving agrochemicals and those from other allergens, and measuring prevalence becomes baffling. Thyssen et al. (2007) and Mirshahpanah and Maibach (2007) determined a median prevalence and summarized the main findings from studies on contact allergy in the general population. These overviews suggest that the weighted average prevalence was 19.5%, based on data collected on all age groups and all countries between 1966 and 2007. The median prevalence was 21.8% in women and 12% in men (Thyssen et al., 2007). Considering the preceding factors, an assessment of the portion of 19.5% due to agricultural agents given the length and scope of this chapter would be impossible. Instead, an introductory presentation of agrochemicals and reported cases serves to familiarize the reader with the occurrence of ACD resulting from these chemicals.

25.4  Chemicals and cases California provides the most complete set of case reports and surveillance information in the United States, although there is also information from Washington, Oregon, Arizona, Texas, Iowa, Wisconsin, South Carolina, and New York. California also includes Standard Industrial Classification (SIC) in most of the reports, which is useful for categorizing pesticides and other chemicals as agricultural and nonagricultural. The main crops cultivated in California include wheat, rice, barley, alfalfa, grain crops, cotton, vegetables, and fruit. This variety allows the study of skin problems related to agents associated with many different types of agricultural production, ranging from mechanized production of cereals (SIC 011) and livestock (SIC 02) to

Chapter | 25  Allergic Contact Dermatitis

the labor-intensive production of fruit (SICs 0171, 0172, and 0175), vegetables (SIC 016), nut crops (SIC0173), and ornamental nursery crops (SIC 018) (Penagos et al., 2001). Indeed, nursery workers are included in a high-risk category for occupational skin disease (O’Malley et al., 1995). O’Malley et al. documented the distribution of 2722 claims for lost-work-time skin conditions reported between 1978 and 1983 according to the chemical’s SIC. The results revealed that employees in horticultural specialties had the highest rate of claims associated with exposure to agricultural chemicals (O’Malley and Mathias, 1988). Workers dispensing or transferring the agrochemicals had the highest rate of skin disease in a report from the Illness Registry and Propargite Outbreak Episodes (Penagos et al., 2001). Some were exposed during direct administration and others by means of hand labor activities, but a large portion of the cases for each involved workers who had directly handled the compound. Most of the skin conditions were associated with handling of grapes due to field residues. Grain and livestock produced lower incidence among the workers. Some common agents associated with irritant and/or allergic contact dermatitis seen in pesticide workers are shown in Table 25.1. Gamsky et al. (1992) measured crop-specific dermatitis prevalence in the farm worker population. The study focused on grape, citrus, and tomato workers in four central California counties. Among six vineyards, two were identified as using sulfur only, and the other four used a variety including triadimefon (Bayleton), methamyl (Lannate), sulfur, and Bacillus thuringiensis (Dipel). Growers reported using chlorpyrifos (Lorsban), formetanate (Carzol), methamyl (Lannate), manganese sulfate, sulfur, copper sulfate, and zinc sulfate on citrus fields. Tomato field agrochemicals included trifluralin (Treflan), napropamide (Deverol), pebu-

Table 25.1  Most Common Chemicals Indicated in Cases of Irritant and/or Allergic Contact Dermatitis in Agrochemical Handlers Pesticide

Classification

Cases (out of 1225)

Pesticide mixtures

Multiple

270

Sodium hypochlorite

Chlorine/chlorine releaser

184

Glyphosate

Herbicide

89

Metam sodium

Fumigant

34

Sulfur

Fungicide

32

Propargite

Insecticide and miticide

31

Methyl bromide

Fumigant

27

From Penagos et al. (2001).

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late (Tillam), glyphosphate (Roundup), sulfur, parathion, maneb (Dithane), ethephon (Ethrel), fenvalerate (Pydrin), dimethoate, and carbaryl (Sevin). Questionnaires and waistup skin examinations were used to determine how many workers had experienced a rash in the past 12 months that lasted 2 or more days. Sixty-two percent of the subjects with contact dermatitis on examination indicated on the questionnaire that they had had a rash lasting more than 2 days in the past 12 months; however, rashes within the past 12 months lasting longer than 2 days were more frequently reported by grape workers than by citrus or tomato workers. Other studies confirmed that the highest rates of pesticiderelated skin conditions among agricultural production SICs were found for grapes (Penagos et al., 2001). Although California only represents one state, its agricultural workforce accounts for approximately one-third of the total U.S. agricultural employment. Thus, the previously discussed reports for California may be useful in estimating national outcomes. Other countries, particularly those in Europe, also provide data on cases involving skin irritation due to agrochemicals. Agrochemicals can be divided into three or four larger categories, including insecticides, fungicides, and herbicides. As previously stated, it is difficult to differentiate which skin cases are ACD or ID in studies lacking patch tests and careful interpretation using relevance grading systems like the one proposed by Lachapelle (1997) to ascertain causation. Organophosphates are insecticides with high octanol/ water partition coefficients that penetrate the skin easily and have a very specialized function. As a group, they cause few skin reactions. Diazinon, or Spectracide, was among the more irritating organophosphates in the guinea pig maximization tests conducted by Matsushita et al. (1985). In the challenge phase, the maximum nonirritant concentration was only 0.5% compared to 2% for naled, malathion, and leptophos. A second insecticide, naled (Dibrom), contains a dibromo­dichloropropanol moiety that loses the bromine quickly in sunlight to form another insecticide called dichlorvos. Naled is used on grapes, and surface residues taken 6 h after application contain approximately 30% dichlorvos (Penagos et al., 2001). Both forms are characterized by a comparatively high vapor pressure, which means they evaporate into the surrounding atmosphere quickly after application. Edmundson and Davies (1967) reported cases of possible contact sensitivity to naled: 24-h provocation tests to naled were positive in 75% of the reported cases but only 12.5% in the control. Organochlorines are insecticides possessing high lipid solubility, which makes them readily absorbable by the skin. Dicofol contains a centrally located hydroxyl group that allows for a more rapid environmental degradation than some of its counterparts. Thirty percent of animals subjected to dicofol by means of the Draize method

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illustrated erythemous reactions that disappeared within 7 days. Some dicofol products are thus labeled as potential sensitizers. Lindane controls insect infestation on commercial crops and some parasites found on cattle. The Buehler guinea pig assay reveals that it is a positive sensitizer. Scabies patients occasionally present with dermatitis after treatment with 1% lindane formulations. The ability of a chemical to act as a fungicide depends on the structure. Electrophilic functional groups allow the inhibition of sterol synthesis enzymes within the fungus and can also interfere with cellular respiration. The properties previously defined for a good fungicide also support reactivity with skin proteins. Thus, fungicides are often implicated as sensitizers and carcinogens. Phthalimido compounds are compounds resembling captan, with differences mainly in the structure of the side chains. Captan is used on grapes, apples, almonds, and other crops. In a repeat insult patch test protocol by Marzulli and Maibach (1973), captan was a potent experimental sensitizer at a concentration of 1.0%. The International and North American Contact Dermatitis Groups consider a concentration of 0.25% in petrolatum to be sufficient in patch tests (Cronin, 1980). Captafol, similar to captan in structure, is used on potatoes, fruits, and grains. Cases of dermatitis were reported by Camarasa (1975) and Lee et al. (1981). Benomyl (Benlate), a fungicide belonging to the carbamate family, is used on fruits, nuts, and vegetables. In high doses, it acts as a cholinesterase inhibitor in humans. Matsushita and Aoyama (1979) declared 2% benomyl to be a strong sensitizing agent. Foliar residues of benomyl have caused ACD (Everhart and Holt, 1982; Hargreave, 1983; Zweig et al., 1983). Another carbamate, maneb, prevents and treats horticultural diseases. Matsushita et al. (1976) found it to be a potent sensitizer in the guinea pig maximization test. Numerous cases are presented by Adams and Manchester (1982) and Piraccini et al. (1991). Paraquat (Gramoxone), along with diquat, is a bipyridyl herbicide. Bipyridyl compounds damage the membranes and cytoplasm of unwanted plants by using a centrally located ammonium ion to generate superoxides during photosynthesis. They are used on potatoes, alfalfa, and soybeans. Although dermatitis resulting from contact with paraquat occurs, specific diagnoses of ACD have not been reported. Phenoxy herbicides are absorbed into the roots and leaves of broadleaf plants and are therefore used on grain crops and turf. 2, 4-Dichlorophenoxyacetic acid (2,4-D) (Weed and Feed Products) is made by adding chlorine to phenol. Although California requires products containing the compound to display a sensitization warning, application of 97% formula did not result in any signs of reaction using the Draize test (Penagos et al., 2001). Contact dermatitis is more commonly described in cases dealing

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with 2,4-D contact, and international cases have indicated reactions lasting for long periods, suggesting a mechanism similar to that for ACD (Sharma and Kaur, 1990). The acetanilides include the herbicides alachlor (Lasso) and metolachlor (Dual and Pennant). Alachlor is applied mainly to corn crops and is registered as a sensitizer. Metolachlor is used on corn, sorghum, legumes, peppers, and tomatoes to control broadleaf weeds. An increase in erythema was noted from induction to challenge in a Buehler guinea pig sensitization study of a 79.9% metolachlor formula (Penagos et al., 2001).

25.5  Diagnosis and treatment of allergic contact dermatitis The appearances of ACD and ID are often similar, and the health care worker must rely on history of contact with the chemical to begin to discern between them (Smith and Hotchkiss, 2001). They both appear as an eczematous rash, which is red, inflamed, dry, and scaly and is sometimes accompanied by discharge. The symptoms vary depending on the sensitivity of the individual and location. ACD is usually ill-defined and extends beyond the site of application of the allergen, in contrast to ID, which often produces lesions with sharp borders (Lachapelle and Maibach, 2009). In addition, the pustules, necrosis, or ulceration associated with ID are rarely seen in ACD, and ID is also accompanied by burning, stinging, pain, and soreness of the skin (Ale and Maibach, 2006). Diagnostic patch testing, combined with algorithms to ascertain their clinical relevance, provides the basis for separating irritant from allergic causality (Ale and Maibach, 1995; Hostynek and Magee, 1997; Lachapelle, 1997; Marrakchi and Maibach, 1994). The diagnosis of ACD is aided by patch testing. Patch test chemicals can be obtained commercially at standardized sensitizing doses in pre-prepared syringes. There are differences in the standard series of chemicals used in different countries, and some have been designed to test individuals working in certain occupations according to the chemicals they are likely to encounter. The chemicals can be applied in various ways. Finn Chambers are round aluminum patches that can be used to test numerous commercial substances. They are small and are applied using Scanpor tape onto the skin with various other tests. Plastic square chambers include models from IQ Square Chamber Chemotechnique, van der Bend, the Haye’s Test Square Chamber, and the allergEAZE Patch Test Chamber. The skin is inspected after 48 and 96 h. Reactions are graded as no reaction (0), doubtful reaction (/), weak (nonvesicular) reaction (), strong (edematous) reaction (), extreme reaction (  ), or irritant (IR) (Smith and Hotchkiss, 2001). Interpretation of the patch tests is

Chapter | 25  Allergic Contact Dermatitis

not a simple positive or negative. The reactions and lack of reactions are often subject to changing variables, such as the type of patch system used, amount of allergen applied, condition of skin area tested, and variations in responses among patients (Lachapelle and Maibach, 2009). Some patch tests come with the allergens already loaded and pre-prepared. The thin-layer rapid use epicutaneous (TRUE) test is pre-prepared and consists of 23 allergens and allergen mixes (Timm-Knudson et al., 2006). The North American Contact Dermatitis Standard Screening Series has 65 allergens and is more commonly used by specialists. Contrary to the TRUE test, it must be assembled and is customized to test the allergens indicated by the patient’s history and exam (Ortiz and Yiannias, 2004). Occupationally and environmentally relevant series have been developed to aid in determining whether rashes are a result of contact allergy to an occupational hazard. Once a specialist decides ACD is the correct diagnosis, avoidance of the causative allergen is usually effective. However, an existing dermatitis will clear more rapidly with application of topical corticosteroids. Caution must be used when prescribing these medications because sensitization to ingredients in the ointment may infrequently exacerbate symptoms or cause recurrence (Smith and Hotchkiss, 2001). Databases of the appropriate patch test concentrations for commonly utilized agrochemicals are limited. Note that optimization of patch test concentration is most efficiently defined in patients/workers who are clinically believed to be sensitized. See O’Malley et al. (1995) for a summary of these data. We have previously attempted to assess the clinical relevance of reports on agricultural chemicals. A major limitation of our analysis relates to the quality of the data on which it is based. Penagos et al. (2001, pp. 89–162) summarize this information. Our caveat is that many of these data are incomplete and based on clinical impression and/or minimal diagnostic patch testing. Patch tests may be truly indicative of disease but may also be false positives and/or negatives. Table 25.2 provides an abbreviated algorithm that permits determination of the likelihood that a positive patch test relates to a clinical entity – ACD (Ale and Maibach, 1995; Hostynek and Magee, 1997; Lachapelle, 1997; Marrakchi and Maibach, 1994). In an ideal situation, all responses would be “yes” in a patient presenting with ACD.

25.6  Prevention Despite the tendency of ACD to subside spontaneously, the symptoms can be emotionally and physically debilitating. Predicting the effects of agrochemicals on the skin is therefore important to both consumers and industry. However, applying experimental findings from human assays to a

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Table 25.2  Morphology Suggestive of Allergic Contact Dermatitis Likely chemical exposure

Yes/no

Positive patch test to appropriate dilution

Yes/no a

Re-patch test to rule out excited skin syndrome

Yes/no

Positive use test (provocative use test) (repeated open application test)

Yes/no

Clinical remission within several weeks of discontinuing exposure

Yes/no

a

Recommended by Ale and Maibach (1995), Hostynek and Magee (1997), Lachapelle (1997), and Marrakchi and Maibach (1994). From Lachapelle and Maibach (2009).

general population may not be accurate due to individual sensitivity, and experiments with human subjects are accompanied by fear of the risk involved. Animal models can also present signs that do not coincide with those seen in humans. Each model is limited by our ability to interpret and extrapolate the features of inflammation to the desired context. Therefore, predicting human responses based on data from nonhuman models requires care (Penagos et al., 2001). Despite drawbacks, animal models have proven useful for predicting hazard identification and dosage necessary to produce ACD. Most studies with animal models utilize the guinea pig to predict sensitization. The guinea pig maximization test took 8 years to develop and was meant to enhance the usefulness of the guinea pig in screening contact allergens. The inspiration was due to physicians’ experience with certain chemicals that had failed to sensitize the guinea pig but were known to cause reactions in humans. This method effectively enhanced the usefulness of the animal in screening allergens by increasing sensitivity of guinea pig skin (Magnussen and Kligman, 1969). The standard outbred Hartley strain of albino guinea pigs is used, and 30–40 animals are sufficient to account for reactions in those tested with the allergen plus a control group. The procedure has three steps: sensitization, induction, and challenge. Sensitization in the animals is induced by using intradermal injections, with some containing allergens incorporated in adjuvant and some with allergens only. The adjuvant, Freund’s common adjuvant, consists of a mixture of paraffin oil and an emulsifier with mycobacteria. Induction occurs in two phases. First, three pairs of injections are made containing Freund’s complete adjuvant only, the test allergen in a vehicle, and a mixture of allergen with Freund’s complete adjuvant. Second, the same area is pretreated 1 week after injections with 10% sodium lauryl sulfate (SLS) in petrolatum 24 h prior to application of a patch containing the allergen. SLS enhances sensitization in the animal by producing local inflammation. The patch

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containing the allergen in petrolatum is applied and secured by an elastic bandage wound around the torso. The challenge test occurs through topical application of the allergen in a Finn Chamber to the flank of the animal. It is evaluated 24 and 48 h after application for any redness or swelling. The allergenicity of the tested chemical can be assessed by evaluating the percentage of guinea pigs sensitized. Hartley guinea pigs are also tested in the split adjuvant experiment. This differs from the previous method in that only topical allergen is needed to initiate sensitization (Maibach and Lowe, 1985). Sensitization is elicited by applying allergen to the anterior flank, once on day 0 and day 2, and covering with filter paper, occlusive tape, and an outside layer of adhesive tape. On day 4, 0.75 ml of Freund’s complete adjuvant is injected intradermally. A third application of allergen is then applied and resealed. The last topical application of allergen occurs on day 7, and all bandages are removed 2 days later. On day 22, the challenge phase begins when a smaller amount of test material is applied to the skin, covered with filter paper, and secured with tape. The next day, the bandages are removed and skin is examined 24 and 48 h later for erythema and edema. All reactions are compared with the control to evaluate whether sensitization occurred. The interpretation of results for an experiment using the previously described models may vary among evaluators. The local lymph node assay offers a more quantitative approach to analysis of the immune response to chemicals. The test begins with application of the substance in question to the ears of mice for 3 days. The concentrations vary each day, and the vehicle chosen depends on the properties and solubility of the agrochemical. A control group is exposed to vehicle only and is used as a comparison. On the fifth day, the animals receive intravenous injections of [3H]methylthymidine and are sacrificed after 5 h. The auricular lymph nodes are removed, and suspensions of lymph node cells are made for each group. The amount of [3H]methylthymidine incorporated is measured using -scintillation. If isotope incorporation is three times greater in the test group than in the control, sensitization has occurred in the mice. The lymph node assay is useful in that it allows for an objective interpretation for sensitization instead of subjective evaluation of dermatitis in the guinea pig (Kimber et al., 1994). Although both tests have improved our ability to predict sensitization potential in humans, no test is an adequate substitute for clinical/field experience. Knowledge of dermatotoxicology and immunology is important in interpreting test results.

Conclusion True frequency for the occurrence of ACD due to agrochemicals is unknown. Inadequate skin health facilities

especially for migrant workers correlate with inaccurate representation of the number of cases. In addition, many examined skin reactions are evaluated without followup patch testing, which increases false reports, and the unavailability of commercially attainable patch tests including agrochemicals creates a necessity for trained health care professionals to assemble tests required for evaluation of sensitivity. There is also a need for public availability for open review of dermatotoxic profiles including predictive irritation and sensitization assay and documented irritation and allergic reactions. Taken together, we believe that further cooperation among government, unions, workers, industry, and academia is necessary to verify frequency, develop better identification of occupational risk, permit development of more efficient personal protections, and develop agrochemicals with lower sensitization and irritant potential.

References Adams, R. M., and Manchester, R. D. (1982). Allergic contact dermatitis to Maneb in a housewife. Contact Dermatitis 8(4), 271. Ale, S., and Maibach, H. I. (1995). Clinical relevance in allergic contact dermatitis. Dermatosen 43, 119–121. Ale, S., and Maibach, H. I. (2006). Irritant contact dermatitis versus allergic contact dermatitis. In “Irritant Dermatitis” (A.-L. Chew and H. I. Maibach, eds.), pp. 11–18. Springer, Berlin. Camarasa, G. (1975). Difolatan dermatitis. Contact Dermatitis 1(2), 127. Cronin, E. (1980). “Contact Dermatitis.” Churchill Livingstone, Edinburgh, UK. Edmundson, W. F., and Davies, J. E. (1967). Occupational dermatitis from naled. A clinical report. Arch. Environ. Health 15, 89–91. Everhart, L. P., and Holt, R. F. (1982). Potential benlate fungicide exposure during mixer/loader operations, crop harvest, and home use. J. Agric. Food Chem. 30(2), 222–227. Gamsky, T., McCurdy, S. A., Wiggins, P., Samuels, S. J., Berman, B., and Shenker, M. B. (1992). Epidemiology of dermatitis among California farm workers. J. Occup. Med. 34(3), 304–310. Hargreave, P. (1983). Benomyl residues on lichtis after post-harvest dipping. Aust. J. Exp. Agric. Anim. Husbandry 23(120), 95–98. Hostynek, J., and Magee, P. S. (1997). Fragrance allergens: Classification and ranking by QSAR. Toxicol. in Vitro 11(4), 377–384. Jacob, S. E., and Steele, T. (2006). Allergic contact dermatitis: Early recognition and diagnosis of important allergens. Dermatol. Nursing Ser. 18(5), 433–439 446. Kimber, I., Dearman, R. J., Scholes, E. W., and Basketter, D. A. (1994). The local lymph node assay: Developments and applications. Toxicology 93, 13–31. Krasteva, M., Kehren, J., Ducluzeau, M. T. et al. (1999). Contact dermatitis I: Pathophysiology of contact sensitivity. Eur. J. Dermatol. 9(1), 65–77. Lachapelle, J. (1997). A proposed relevance scoring system for positive allergic patch test reactions: Practical implications and limitations. Contact Dermatitis 36, 39–43. Lachapelle, J.-M., and Maibach, H. I. (2009). The standard series of patch tests. In “Patch Testing and Prick Testing: A Practical Guide”, pp. 71–82. Springer-Verlag, Berlin. Lee, S., Cinn, Y., Chang, W., and Kim, J. (1981). A study on hypersensitivity of Korean farmers to various agrochemicals: 1. Determination

Chapter | 25  Allergic Contact Dermatitis

of concentration for patch test of fruit-tree agrochemicals and hypersensitivity of orange orchard farmers in Che-ju Do, Korea. Seoul J. Med. 22(1), 137–142. Magnussen, B., and Kligman, A. (1969). The identification of contact allergens by animal assay. The guinea pig maximization test. J. Invest. Dermatol. 52(3), 268–276. Maibach, H., and Lowe, N. J. (eds.) (1985). “Models in Dermatology”, Vol. 1. Karger, Basel, Switzerland. Marrakchi, S., and Maibach, H. I. (1994). What is occupational contact dermatitis? An operational definition. Dermatol. Clin. 12(3), 477–484. Marzulli, F., and Maibach, H. (1973). Antimicrobials: Experimental contact sensitization in man. J. Soc. Cosmetic Chemists Jpn. 24(7), 399–421. Matsushita, T., and Aoyama, K. (1979). Examination on cross sensitivity between benomyl and other major pesticides in hypersensitive contact dermatitis. Nippon Noson Igakkai Zasshi 28, 464-465. Matsushita, T., Arimatsu, Y., and Nomura, S. (1976). Experimental study on contact dermatitis caused by dithiocarbamates maneb, mancozeb, zineb, and their related compounds. Int. Arch. Occup. Environ. Health 37(3), 169–178. Matsushita, T., Aoyama, K., Yoshimi, K., Fujita, Y., and Ueda, A. (1985). Allergic contact dermatitis from organophosphorous insecticides. Ind. Health 23(2), 145–153. Mirshahpanah, P., and Maibach, H. (2007). Relationship of patch test positivity in a general versus an eczema population. Contact Dermatitis 56, 125–130. Muller, G., Knop, J., and Enk, A. H. (1996). Is cytokine expression responsible for differences between allergens and irritants? Contact Dermatitis 7, 177–184.

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O’Malley, M., and Mathias, C. G. T. (1988). Distribution of lost-worktime claims for skin disease in California agriculture: 1978-1983. Am. J. Ind. Med. 14(6), 715–720. O’Malley, M., Rodriguez, P., and Maibach, H. I. (1995). Pesticide patch testing: California nursery workers and controls. Contact Dermatitis 32(1), 61–63. Ortiz, K. J., and Yiannias, J. A. (2004). Contact dermatitis to cosmetics, fragrances, and botanicals. Dermatol. Ther. 17, 264–271. Penagos, H., O’Malley, M., and Maibach, H. (eds) (2001). “Pesticide Dermatoses”. CRC Press, Boca Raton, FL. Piraccini, B. M., Cameli, N., Peluso, A. M., and Tardio, M. (1991). A case of allergic contact dermatitis due to the pesticide maneb. Contact Dermatitis 24(5), 381–382. Saint-Mezard, P., Bérard, F., Dubois, B. et al. (2004). Allergic contact dermatitis. Eur. J. Dermatol. 14(5), 284–295. Sharma, V. K., and Kaur, S. (1990). Contact sensitization by pesticides in farmers. Contact Dermatitis 23(2), 77–80. Smith, C. K., and Hotchkiss, S. A. M. (2001). “Allergic Contact Dermatitis”. Taylor & Francis, London. Thyssen, J., Uter, W., Schnuch, A., Linneberg, A., and Johansen, J. D. (2007). The epidemiology of contact allergy in the general population – Prevalence and main findings. Contact Dermatitis 57, 287–299. Timm-Knudson, V. L., Johnson, J. S., Ortiz, K., and Yiannias, J. (2006). Allergic contact dermatitis to preservatives. Dermatol. Nursing 18(2), 130–136. Zweig, G., Gao, R. Y., and Popendorf, W. (1983). Simultaneous dermal exposure to captan and benomyl by strawberry harvesters. J. Agric. Food Chem. 31(5), 1109–1113.

Chapter 26

Pesticides and Contact Urticaria Syndrome Iris S. Ale1 and Howard I. Maibach2 1 2

Republic University of Uruguay, Montevideo, Uruguay University of California, San Francisco, California

26.1  Definition The contact urticaria syndrome (CUS; immediate contact reactions), first defined as a biological entity in 1975 by Maibach and Johnson, comprises a wide range of inflammatory reactions that habitually appear within minutes after cutaneous or mucosal contact with the causal agent and disappear within 24 h after contact. However, delayedonset reactions appearing several hours after exposure are sometimes observed (von Krogh and Maibach, 1981). The term “syndrome” clearly illustrates the biological and clinical polymorphism of this entity, which may be either localized or generalized and may involve organs other than the skin, such as the respiratory or the gastrointestinal tract as well as the vascular system. CUS usually displays a broad spectrum of clinical manifestations ranging from mild erythema and/or itching or burning to severe anaphylactic reactions and even death (Odom and Maibach, 1976; von Krogh and Maibach, 1982, 1984). Numerous cases of CUS continue to be reported, and the list of etiologic agents cons­tantly increases (Amin et al., 1997; Burdick and Mathias, 1985; Harvell et al., 1992; Lahti and Maibach, 1987, 1991). It has been reported following skin contact with a multitude of substances ranging from simple chemicals to macromolecules. Its prevalence among the general population is unknown, but it may be a relatively common and under-recognized phenomenon.

26.2  Clinical signs and symptoms CUS can be categorized according to the clinical signs and symptoms as well as overall severity (Table 26.1) (Odom and Maibach, 1976; von Krogh and Maibach, 1981). In the invisible contact urticaria, only subjective symptoms (itching, tingling, or burning) without any objective sign or just Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

Table 26.1  The Contact Urticaria Syndrome: Staging by Symptomatology Cutaneous reactions only Stage 1

Localized urticaria Dermatitis Nonspecific symptoms (itching, tingling, burning, etc.)

Stage 2

Generalized urticaria

Cutaneous and extracutaneous reactions Stage 3

Rhinoconjunctivitis Orolaryngeal symptoms Bronchial asthma Gastrointestinal symptoms

Stage 4

Anaphylactic symptoms

Adapted from von Krogh and Maibach (1982).

a mild erythema occur. These reactions are often seen from cosmetics and from fruits and vegetables. A localized whealand-flare reaction following external contact with a substance is the prototype of contact urticaria, whereas generalized urticaria after a local contact is less common. Extracutaneous symptoms may also occur as part of a more severe reaction and may include rhinoconjunctivitis, asthmatic attack, and orolaryngeal or gastrointestinal dysfunctions. Finally, anaphylaxis may occur as the most serious manifestation of CUS. Contact urticaria (CU) usually clears spontaneously, usually in a few hours; repeated exposure may lead to dermatitis (eczema) (Ale and Maibach, 2000a; Maibach, 1976). 677

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In addition, CU may be associated with allergic contact dermatitis (type IV hypersensitivity). von Krogh and Maibach (1982) suggested that the term “contact dermatitis of immediate and delayed type” should be used for patients exhibiting both types of reactions to the same agent. Not just CU may produce dermatitis; immediate contact reactions aggravating chronic dermatitis have been reported (Ale and Maibach, 2000a; Amin et al., 1997; Harvell et al., 1992; von Krogh and Maibach, 1984). A previous irritant contact dermatitis produced by the working environment may predispose not only to allergic contact dermatitis but also to immediate contact reactions. A defective skin barrier function might facilitate the penetration of macromolecules such as protein allergens that have been proven to be responsible for most of the immediate contact-type reactions (Ale and Maibach, 2000a; Amin et al., 1997; Cromwell, 1997; Maibach, 1976).

26.3  Etiology and mechanisms Contact urticaria is classified as nonimmunological or immunological according to the underlying mechanism. A third category exists for reactions with mixed features or undetermined pathomechanisms (Amin et al., 1997; Cromwell, 1997; Harvell et al., 1992).

26.3.1  Immunologic Contact Urticaria Immunological contact urticaria (ICU) is a type I hypersensitivity reaction, mediated by allergen-specific IgE in a previously sensitized individual. The consequences are potentially more serious than for nonimmunological contact urticaria (NICU) because reactions may not remain localized to the area of contact, and generalized urticaria, or even involvement of internal organs such as the respiratory or gastrointestinal tract, may follow, leading to anaphylactic shock (Amaro and Gossens, 2007; Hannuksela, 1997; Harvell et al., 1994; Wakelin, 2001). Even very small amounts of allergens can induce a reaction, and only selected (sensitized) individuals among the exposed persons are affected. ICU develops preferentially in persons with an atopic predisposition or atopic diseases (Harvell et al., 1994; Wakelin, 2001). The list of substances inducing proven or probable antigen-specific, IgE-mediated CU is considerable. Next to latex, foods are the most frequent eliciting agents. Plants, animal products, drugs, cosmetics, and industrial products may also be elicitors of ICU. Agents responsible for ICU reactions are mostly proteins, but other substances, such as low-molecular-weight agents (haptens), may cause IgE-mediated type I allergic reactions (Amaro and Gossens, 2007; Hannuksela, 1997; Harvell et al., 1994; Wakelin, 2001). The elicitating substance, pene­ trating through the skin or mucosal membrane, will react with two adjacent IgE molecules bound to the cell membrane of the mast cell (Sutton and Gould, 1997). Within minutes,

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h­ istamine, exoglycosidases, neutral proteases, and proteoglycans are released from the mast cells, resulting in an immediate skin response. This reaction comprises local edema (wheal) and erythema dependent on reflex neural stimulation (flare). Massive amounts of these active substances lead to anaphylaxis. The allergen-IgE reaction also leads to synthesis of leukotrienes, prostaglandins, and platelet-activating factors in the cell membranes of the activated mast cells. These mediators acting on endothelial cells increase vascular permeability. No single mediator accounts for all of the components of the IgE-dependent response; many substances have similar actions resulting in additive effects. In addition, IgE molecules bind to the high-affinity receptors on antigen-presenting dendritic cells (APCs) (Ale and Maibach, 2000a; Barker et al., 1988; Bruynzeel-Koomen, 1986). APCs can present protein allergens to the T helper 2, inducing a delayed-type hypersensitivity reaction resulting in eczematous lesions (Barker et al., 1988; Bruynzeel-Koomen, 1986; Nethercott et al., 1984).

26.3.2  Nonimmunologic Contact Urticaria In NICU, the most frequent type of contact urticaria, no previous sensitization has occurred and the agents will produce CU in most individuals if contact time and concentration are sufficient (Lahti, 1980). The NICU reaction is often redness without edema rather than a real wheal-and-flare reaction. The appearance of clinical signs depends mainly on the duration of exposure, the concentration of the contactant, and other factors, such as rubbing or scratching. With few exceptions, the reaction remains localized to sites of contact and rarely evokes systemic manifestations (Lahti, 1980; Odom and Maibach, 1976; von Krogh and Maibach, 1981, 1982, 1984). Substances capable of producing NICU are usually low-molecular-weight chemicals that easily cross the barrier of the skin. Many of the chemical substances involved are used as flavorings, fragrances, and preservatives in the cosmetic, pharmaceutical, and food industries. Other responsible agents include plants, animals, or industrial and laboratory chemicals including insecticides. The pathogenesis of NICU is not fully understood, but it appears to involve the release of vasogenic mediators without participation of specific immune mechanisms (Barker et al., 1988). Many of the eliciting agents in this category may induce urticarial skin reactions via several mechanisms: direct release of histamine and other mediators from mast cells; direct effect on dermal vessels; release of vasoactive amines, acetylcholine, leukotrienes, and prostaglandins; and others (Cromwell, 1997; Hannuksela, 1997; Harvell et al., 1994; Lahti, 1980; Lahti et al., 1983; Nethercott et al., 1984).

26.3.3  Uncertain Mechanism A third category of CU consists of cases of uncertain mechanism. In some instances, the reaction resembles that

Chapter | 26  Pesticides and Contact Urticaria Syndrome

of ICU, but no specific IgE can be demonstrated in the patient’s serum or in the tissues. It is possible that there are other immunologic mechanisms in addition to the IgEmediated ones (von Krogh and Maibach, 1983).

26.4  Animal and human assays Predictive assays for evaluating the capacity of substances to produce NICU have been developed. Lahti and Maibach (1984) developed the guinea pig ear-swelling test as a quantitative model to screen human NICU agents. Hartley guinea pigs were challenged by applying 50 l of various concentrations of human NICU agents – benzoic acid (BA), sorbic acid (SA), cinnamic acid (CA), cinnamaldehyde (CAL), methyl nicotinate (MN), and dimethyl sulfoxide (DMSO)�������������������������������������������������������� – ��������������������������������������������������� in absolute ethyl alcohol to both sides of the earlobe. The thickness of the ear was measured with a string micrometer before application and then every 15 min for 2 h after application. The swelling response was dependent on the concentration of the elicitating substance. Maximal increase in ear thickness was produced within 30–40 min by 20% BA, 10% SA, 15% CA, 5.0% CAL, 0.2% MN, and 100% DMSO. All responses were dose dependent. A long refractory period up to 16 days after application was observed with the different substances (Lahti and Maibach, 1985a). Guinea pig body skin reacts with quick-appearing isolated erythema to CAL, MN, and DMSO, but BA, SA, and CA did not cause any reactions. Analogous reactions can be elicited in the earlobes of other animal species. CAL and DMSO induce a swelling reaction in the guinea pig, rat, and mouse. On the contrary, BA, SA, CA, and MN produce no response in the rat or mouse, but the guinea pig ear reacts to all of them (Lahti and Maibach, 1985b). This suggests that either there are several mechanisms of NICU or differences are due to relative sensitivity of the species to the mediators. The guinea pig ear swelling test remains the best quantitative animal method available for screening human NICU agents (Lahti and Maibach, 1985b). Materials can also be screened for NICU in man (Gollhausen and Kligman, 1985; von Krogh and Maibach, 1982). A small amount of the substance is applied to a marked site in the forehead and the vehicle is applied to a parallel site. The areas are evaluated approximately 20–30 min after application for erythema and/or edema. Currently, there is no appropriate model available for ICU.

26.5  Diagnosis of CUS When assessing CUS, a detailed anamnesis, physical examination, and diagnostic testing should be performed. Clinical history must investigate any occurrence of immediate reactions – whether limited to the skin or not – as well as all suspicious occupational and nonoccupational exposures. CUS has a large heterogeneity of clinical ­ manifestations;

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therefore, patients may disclose a variety of symptoms. Sometimes patients experience only subjective symptoms such as itching, burning, or tingling, which can be easily disregarded if the physician is not alert to the possibility of CUS. In addition, CUS has to be considered in patients developing other immediate-type reactions, such as bronchial asthma or rhinoconjunctivitis. In vivo tests for immediate IgE-mediated allergy are of paramount importance in the evaluation of CUS. Guidelines for skin testing in CUS have been suggested by von Krogh and Maibach (1981) and Odom and Maibach (1976). Tests should first be performed on healthy skin, preferably in open application. The substance should be applied as is or, if necessary, diluted in an excipient. Open application tests may produce negative results unless the substance is applied on damaged or eczematous skin. Therefore, if an open test is negative, it should be repeated on a slightly affected (or previously affected) area. Sometimes a rubbing test (gentle rubbing with the material) on intact or lesional skin might be indicated. Scratch and scratch-patch testing carry a higher risk of false-positive reactions, and the latter lacks sensitivity compared with prick testing. Prick tests with fresh material or commercial reagents are the gold standard for immediate hypersensitivity reactions and should be performed if open application tests with a suspicious substance are negative. Following the recommended order is important for minimizing the occurrence of hazardous extracutaneous or anaphylactic reactions (Ale and Maibach, 2000b). Life-threatening reactions when performing skin tests have been reported (Haustein, 1976; Kosáková, 1977; Maucher, 1972). Therefore, skin tests should be performed only if resuscitation equipment and trained personnel are available. Immunologic-type agents may also be studied by in vitro tests, such as radioallergosorbent tests. However, many of the substances that elicit ICU have not been identified. In vitro tests can also be used to assess the relationship with a possible IgE-mediated mechanism, namely determination of histamine release from peripheral basophils or platelet cytotoxicity test, which explores the presence on platelets of specific IgE antibodies bound to the low-affinity receptor for IgE.

26.6  CUS induced by pesticides Agriculture has consistently had the highest rates and numbers of occupational skin diseases reported in the United States (Mathias and Morrison, 1988). Occupational exposure to pesticides in agricultural workers may occur while spraying the pesticide on fields, loading and mixing, spreading pesticide-preserved seeds, and harvesting previously sprayed crops, as well as while cleaning tools and disposing of empty containers (O’Malley, 2001). Nonoccupational exposure to pesticide products and their residues may also occur from household use and also through soil, water, and

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food supply (National Center for Environmental Health, 2005). The acute toxic events related to pesticide exposure and issues of long-term carcinogenicity have been the focus of most toxicologic reports (Abrams et al., 1991). Pesticides are responsible for acute poisoning as well as for longterm health effects, including cancer and adverse effects on reproduction. In addition, there are many adverse effects not immediately related to the toxic potential of pesticides, the majority of which are skin diseases. Most pesticide-related dermatoses are contact dermatitis, either allergic or irritant. However, other less frequent dermatoses, such as contact urticaria, erythema multiforme, photoallergy, porphyria cutanea tarda, and chloracne, have been described (Abrams et al., 1991; Cellini and Offidani, 1994; Cole et al., 1997; Guo et al., 1996; Lisi, 1992; Mark et al., 1999; O’Malley, 1997; Paulsen, 1998; Sharma and Kaur, 1990; Spiewak, 2001). CUS was described by Maibach and Johnson in 1975 à propos of a 35-year-old woman who had CU due to the insect repellent diethyltoluamide (DEET; Maibach and Johnson, 1975). The experimental data suggested that it was due to an immunologic response. Later, two other cases of CUS due to DEET were reported. Vozmediano et al. (2000) reported a 16-year-old female with CU related to a commercial product containing 20% DEET. She had positive open application tests to the commercial product as well as to DEET at 100 and 1%. Mayenburg and Rakoski (1983) described a 4-year-old girl who developed CUS to DEET. All cases were supposed to be of immunologic nature (ICU). The fungicide chlorothalonil (tetrachloroisothalonitrile) is used extensively for the protection of various horticultural and fruit crops against fungal infections. In banana plantations, it is used in fumigations by airplanes. It is also used as fungicide in wood preservation and as a preservative in paints. It can induce CUS and anaphylactic symptoms, allergic and irritant contact dermatitis, erythema dyschromicum perstans, and folliculitis, mainly in agricultural workers (Boman et al., 2000; Dannaker and Maibach, 1990; Dannaker et al., 1993; Lensen et al., 2007; Penagos et al., 1996). Dannaker and Maibach (1990) and Dannaker et al. (1993) reported a nursery worker who developed facial erythema and edema accompanied by immediate respiratory symptoms, such as a tight chest and throat after entering the nursery greenhouse, without direct skin contact with chlorothalonil. Testing on normal skin with chlorothalonil (0.01% aqueous) resulted in an anaphylactic reaction. Pyrethrum, an insecticide extract derived from dried, ground flower heads of Chrysanthemum cinerariifolium, has been reported to induce CU, anaphylactic reactions, and respiratory disease in humans (Carlson and Villaveces, 1977; Culver et al., 1988; Newton and Breslin, 1983; Potter et al., 1991; Wagner, 1994). An evidence-based review of the literature by Franzosa et al. (2007) that included more than 250 articles published during the past century concluded that only 3 reports suggested possible skin manifestations of CU-like

symptoms upon exposure to pyrethrum and/or pyrethrumcontaining products, and none fulfilled the criteria for ICU. Current extraction techniques in refined pyrethrum probably determine the absence of significant proteins speculated to induce type I hypersensitivity (Franzosa et al., 2007). Sodium pentachlorophenate, widely used as a wood preservative, herbicide, fungicide, and weed killer, was reported to cause generalized urticaria and angioedema after cutaneous contact (CUS) in a 39-year-old man who had had occupational exposure to pentachlorophenate for 9 years (Kentor, 1986). However, the skin testing was not controlled, and its interpretation was unclear because the reaction was elicited more than 4 h after the experimental exposure and was not reproducible at the contact site. Zinc diethyldithiocarbamate (ZDC) is a fungicide and insecticide that is also utilized in the process of rubber manufacture. ZDC and chemically related thiocarbamates have been reported as one of the most frequent causes of allergic contact dermatitis in farmers (Sharma and Kaur, 1990). These chemicals, which have been reported as a cause of CU in workers involved in rubber manufacture (Helander and Makela, 1983), have not been reported to cause CU when exposure occurs as a pesticide.

Conclusion Assessing and managing the occupational health risks posed by the use of pesticides in agriculture is a complex but essential task for occupational health specialists and toxicologists. The experience of many countries has shown that prevention of health risk caused by pesticides is technically feasible and economically rewarding for individuals and the entire community. The ability of pesticides to produce immediate contact reactions must be specifically investigated, as has been the case for delayed-type contact sensitizers. Studies on the mechanisms of immediate contact reactions, development of appropriate models, and standardization of diagnostic tests constitute a challenge for further research. Taken together, we suspect that increased awareness of the signs and symptoms of CUS will lead to increased frequency of diagnosis, which is especially important for ICU. Product information formulated by the registrant would help the health care community (and the worker/patient) by including what is known (and not known) of the potential for developing immunologic and nonimmunologic contact urticaria.

References Abrams, K., Hogan, D. J., and Maibach, H. I. (1991). Pesticide-related dermatoses in agricultural workers. Occup. Med. 3, 463–492. Ale, S. I., and Maibach, H. I. (2000a). Contact urticaria and hand eczema. In “Hand Eczema” (T. Menne and H. I. Maibach, eds.), 2nd ed., pp. 387–405. CRC Press, Boca Raton, FL.

Chapter | 26  Pesticides and Contact Urticaria Syndrome

Ale, S. I., and Maibach, H. I. (2000b). Occupational contact urticaria. In “Handbook of Occupational Dermatology” (L. Kanerva, P. Elsner, J. E. Wahlberg, and H. I. Maibach, eds.), pp. 200–216. Springer, New York. Amaro, C., and Gossens, A. (2007). Immunological occupational contact urticaria and contact dermatitis from proteins: a review. Contact Dermatitis 58, 67–75. Amin, S., Tanglertsampan, C., and Maibach, H. I. (1997). Contact urticaria syndrome: 1997. Am. J. Contact Dermatitis 8, 15–19. Barker, J. N. W. N., Alegre, V. A., and MacDonald, D. M. (1988). Surfacebound immunoglobulin E on antigen presentating cells in cutaneous tissue of atopic dermatitis. J. Invest. Dermatol. 90, 117–121. Boman, A., Montelius, J., Rissanen, R. L., and Lidén, C. (2000). Sensitizing potential of chlorothalonil in the guinea pig and the mouse. Contact Dermatitis 43, 273–279. Bruynzeel-Koomen, C. (1986). IgE on Langerhans cells: new insights into the pathogenesis of atopic dermatitis. Dermatologica 172, 181–183. Burdick, A. E., and Mathias, C. G. T. (1985). The contact urticaria syndrome. Dermatol. Clin. 3, 71–84. Carlson, J. E., and Villaveces, J. W. (1977). Hypersensitivity pneumonitis due to pyrethrum. JAMA 237, 1718–1719. Cellini, A., and Offidani, A. (1994). An epidemiological study on cutaneous diseases of agricultural workers authorized to use pesticides. Dermatology 189, 129–132. Cole, D. C., Carpio, F., Math, J. J., and Leon, N. (1997). Dermatitis in Ecuadorian farm workers. Contact Dermatitis 37, 1–8. Cromwell, O. (1997). Biochemistry of the allergens. In “Allergy and Allergic Diseases” (A. B. Kay, ed.), Vol. 2, pp. 797–810. Blackwell, Oxford. Culver, C. A., Malina, J. J., and Talbert, R. L. (1988). Probable anaphylactoid reaction to pyrethrin pediculocide shampoo. Clin. Pharm. 7, 846–849. Dannaker, C. J., and Maibach, H. I. (1990). Contact urticaria and anaphylaxis to chlorothalonil [Abstract]. Am. J. Contact Dermatitis 1, 65. Dannaker, C. J., Maibach, H. I., and O’Malley, M. (1993). Contact urticaria and anaphylaxis to the fungicide chlorothalonil. Cutis 52, 312–315. Franzosa, J. A., Osimitz, T. G., and Maibach, H. I. (2007). Cutaneous contact urticaria to pyrethrum – Real? Common? or Not Documented? An evidence-based approach. Cutaneous Ocular Toxicol. 26, 57–72. Gollhausen, R., and Kligman, A. M. (1985). Human assay for identifying substances which induce non-allergic contact urticaria: the NICU test. Contact Dermatitis 13, 98–106. Guo, Y. L., Wang, B. J., Lee, C. C., and Wang, J. D. (1996). Prevalence of dermatoses and skin sensitization associated with use of pesticides in fruit farmers of southern Taiwan. Occup. Environ. Med. 53, 427–431. Hannuksela, M. (1997). Mechanisms in contact urticaria. Clin. Dermatol. 15, 619–622. Harvell, J., Bason, M., and Maibach, H. I. (1992). Contact urticaria (immediate reaction syndrome). Clin. Rev. Allergy 10, 303–323. Harvell, J., Bason, M., and Maibach, H. (1994). Contact urticaria and its mechanisms. Food Chem. Toxicol. 32, 103–112. Haustein, U. F. (1976). Anaphylactic shock and contact urticaria after the patch test with professional allergens. Allergy Immunol. 22, 349. Helander, I., and Makela, A. (1983). Contact urticaria to zinc diethydithiocarbamate (ZDC). Contact Dermatitis 4, 327–328. Kentor, P. M. (1986). Urticaria from contact with pentachlorophenate. JAMA 256, 3350. Kosáková, M. (1977). Sub-Schock bei der Epikutanprobe mit Chloramphenicol. Berfsdermatosen 25, 134–135. Lahti, A. (1980). Nonimmunologic contact urticaria. Acta Derm. Venereol. (Stockh.) 60(Suppl. 1), 50.

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Lahti, A., and Maibach, H. I. (1984). An animal model for nonimmunologic contact urticaria. Toxicol. Appl. Pharmacol. 76, 219–224. Lahti, A., and Maibach, H. I. (1985a). Long refractory period after one application of nonimmunologic contact urticaria agents to the guinea pig ear. J. Am. Acad. Dermatol. 13, 585–589. Lahti, A., and Maibach, H. I. (1985b). Species specificity of nonimmunologic contact urticaria: guinea pig, rat, and mouse. J. Am. Acad. Dermatol. 13, 66–69. Lahti, A., and Maibach, H. I. (1987). Immediate contact reactions: Contact urticaria syndrome. Semin. Dermatol. 6, 313–320. Lahti, A., and Maibach, H. I. (1991). Immediate contact reactions: Contact urticaria and contact urticaria syndrome. In “Dermatotoxicology” (F. N. Marzulli and H. I. Maibach, eds.), pp. 473–495. Hemisphere, New York. Lahti, A., Oikarinen, A., Ylikorkala, O. et al. (1983). Prostaglandins in contact urticaria induced by benzoic acid. Acta Derm. Venereol. (Stockh.) 63, 425–427. Lensen, G., Jungbauer, F., Gonçalo, M., and Coenraads, P. J. (2007). Airborne irritant contact dermatitis and conjunctivitis after occupational exposure to chlorothalonil in textiles. Contact Dermatitis 57, 181–186. Lisi, P. (1992). Pesticides in occupational contact dermatitis. Clin. Dermatol. 10, 175–184. Maibach, H. I. (1976). Immediate hypersensitivity in hand dermatitis. Arch. Dermatol. 112, 1289–1291. Maibach, H. I., and Johnson, H. L. (1975). Contact urticaria syndrome. Contact urticaria to diethyltoluamide (immediate type hypersensitivity). Arch. Dermatol. 111, 726–730. Mark, K. A., Brancaccio, R. R., Soter, N. A., and Cohen, D. E. (1999). Allergic contact and photoallergic contact dermatitis to plant and pesticide allergens. Arch. Dermatol. 135(1), 67–70. Mathias, C. G. T., and Morrison, J. H. (1988). Occupational skin diseases, United States. Results from the Bureau of Labor Statistics Annual Survey of Occupational Injuries and Illnesses, 1973 through 1984. Arch. Dermatol. 10, 1519–1524. Maucher, O. M. (1972). Anaphylaktische Reaktionen beim Epicutantest. Hautarzt 23, 139–140. Mayenburg, J., and Rakoski, J. (1983). Contact urticaria to diethyltolu­ amide. Contact Dermatitis 9, 171. National Center for Environmental Health, Division of Laboratory Sciences (2005). “National Report on Human Exposure to Environmental Chemicals”, Pub. No. 05-0570. Centers for Disease Control and Prevention, Atlanta, GA. Nethercott, J. R., Lawrence, M. J., Roy, A. M., and Gibson, B. L. (1984). Airborne contact urticaria due to sodium benzoate in a pharmaceutical manufacturing plant. J. Occup. Med. 26, 734–736. Newton, J. G., and Breslin, A. B. (1983). Asthmatic reactions to commonly used aerosol insect killer. Med. J. Aust. 1, 378–380. Odom, R. B., and Maibach, H. I. (1976). Contact urticaria: A different contact dermatitis. Cutis 18, 672–675. O’Malley, M. A. (1997). Skin reactions to pesticides. Occup. Med. 12, 327–345. O’Malley, M. A. (2001). Work activities and patterns of skin exposure to pesticides. In “Pesticide Dermatoses” (H. Penagos, M. O’Malley, and H. I. Maibach, eds.), pp. 55–66. CRC Press, Boca Raton, FL. Paulsen, E. (1998). Occupational dermatitis in Danish gardeners and greenhouse workers (II). Etiological factors. Contact Dermatitis 38, 14–19. Penagos, H., Jimenez, V., Fallas, V., O’Malley, M., and Maibach, H. I. (1996). Chlorothalonil, a possible cause of erythema dyschromicum perstans (ashy dermatitis). Contact Dermatitis 35, 214–218.

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Potter, P. C., Berman, D., Toerien, A., Malherbe, D., and Weinberg, E. G. (1991). Clinical significance of aero-allergen identification in the western Cape. S. Afr. Med. J. 79, 80–84. Sharma, V. K., and Kaur, S. (1990). Contact sensitization by pesticides in farmers. Contact Dermatitis 2, 77–80. Spiewak, R. (2001). Pesticides as a cause of occupational skin diseases in farmers. Ann. Agric. Environ. Med. 8(1), 1–5. Sutton, B. J., and Gould, H. J. (1997). IgE and IgE receptors. In “Allergy and Allergic Diseases” (A. B. Kay, ed.), Vol. 2, pp. 797–810. Blackwell, Oxford. von Krogh, G., and Maibach, H. I. (1981). The contact urticaria syndrome – An update review. J. Am. Acad. Dermatol. 5, 328–342. von Krogh, G., and Maibach, H. I. (1982). The contact urticaria syndrome. Semin. Dermatol. 1, 59–66.

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von Krogh, G., and Maibach, H. I. (1983). Contact urticaria. In “Occupational Skin Disease” (R. M. Adam, ed.), pp. 58–69. Grune & Stratton, New York. von Krogh, G., and Maibach, H. I. (1984). The contact urticaria syndrome and associated disease entities. In “Dermatology” (S. L. Moschella, D. M. Pillsbury, and H. J. Hurley, eds.), 2nd ed. Saunders, Philadelphia. Vozmediano, J. M., Armario, J., and Gonzalez-Cabrerizo, A. (2000). Immunologic contact urticaria from diethyltoluamide. Int. J. Dermatol. 39(11), 876–877. Wagner, S. L. (1994). Allergy from pyrethrin or pyrethroid insecticide. J. Agromed. 1, 39–45. Wakelin, S. H. (2001). Contact urticaria. Clin. Exp. Dermatol. 26, 132–136.

Chapter 27

Agricultural Chemical Percutaneous Absorption and Decontamination Heidi P. Chan, Hongbo Zhai, Ronald C. Wester, and Howard I. Maibach University of California, San Francisco, California

27.1  Introduction Percutaneous absorption is a primary focal point for dermatotoxicology and dermatopharmacology. Local and systemic toxicity depend on a chemical penetrating the skin. The skin is a barrier to absorption and a primary route to the systemic circulation. The skin’s barrier properties are impressive. Fluids and precious chemicals are reasonably retained within the body; at the same time, many foreign chemicals are inhibited from entering the systemic circulation. Even with these impressive barrier properties, the skin is a primary body organ that contacts the environment and is a route by which many chemicals enter the body. Some chemicals applied to skin have proved to be toxic, including agricultural chemicals, which in actuality are designed poisons. Table 27.1 summarizes knowledge of parathion gained during the past 30 years. Absorption of parathion was established for human skin contact, but other species similarly absorb the compound. Mathematical models based on quantitative structure–activity relationships now can estimate a human skin permeability coefficient but the accuracy of the predicted coefficient is not fully validated for humans in vivo (Farahmand and Maibach, 2008). Skin absorption amounts combined with toxicity data can predict potential human health hazard. Figure 27.1 shows human systemic parathion absorption from dermal exposure. Parathion is predicted to be lethal not only for total systemic absorption but also for exposure to limited regions. The LD50 used for parathion is 14 mg/kg. Given a body weight of 70 kg, systemic absorption of 980 mg might result in 50% mortality. Thus,

Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

parathion lethal toxicity levels can be reached at 8-h and longer exposures, and unfortunately this was validated in the agricultural fields of California and elsewhere, leading to its ban in California (Wiles et al., 1999). On October 31, 2003, the U.S. Environmental Protection Agency (EPA) and Cheminova (a chemical manufacturer) agreed to discontinue the use of ethyl parathion in corn, alfalfa, barley, canola, sorghum, sunflower, wheat, and soybean (U.S. EPA, 2003).

27.2  Percutaneous absorption methodology 27.2.1  Absolute Topical Bioavailability A reliable way to determine the absolute bioavailability of a topically applied compound is to measure the compound by specific assay in blood or urine after topical and intravenous administration. This is often difficult to do in plasma because concentrations after topical administration are often low. However, with advances in analytical methodology resulting in more sensitive assays, estimates of absolute topical bioavailability have become increasingly available (Wester and Maibach, 1999).

27.2.2  Radioactivity in Excreta Percutaneous absorption in vivo is usually determined by the indirect method of measuring the chemical, metabolite, or radioactivity in excreta after topical application of labeled compound. Radioactive and stable isotope methods

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Table 27.1  Summary of Parathion Percutaneous Absorption Parathion O,O-diethyl O-(4-nitrophenyl) phosphorothioate Other names: ethylparathione. parathion-ethyl CAS: 56-38-2; mol. wt. 291.26 Molecular formula: C10HI4NO5PS Nonsystemic contact and stomach-acting insecticide and acaricide with some fumigant action Nonphytotoxic except to some ornamentals and under certain weather conditions; absorption takes place readily through any portal; fatal human poisoning has followed skin exposure Skin absorption (species)a Human (forearm): solvent, acetone; 10%; 5 days (excretion analysis)b Mouse (dermal): no solvent; 1.4%, 1 h (excretion analysis) Mouse (dermal): solvent, acetone; 32%; days (patch)c Human: solvent, acetone; forearm 8.6%, palm 11.8%, foot 13.5%, abdomen 18.5%, dorsum of hand 21.0%, forehead 36.3%, axilla 64.0%, jaw 33.9%, fossa cubitalis 28.4%, scalp 32.1%, ear canal 46.6%, scrotum 10 1.6%d Frog (dermal): solvent, acetone; 33%; 1 h (patch)e Quail (dermal): solvent, acetone; 40%; 1 h (patch)e Rat (dermal): solvent, acetone; 59%; 1 h (patch)f Roach (dermal): solvent, acetone; 14%; 1 h (patch)e Skin absorption (mathematical model) kp (cm/h) Log P(Ko/w) 1.59  102 3.83 Based on the formula logkp  2.74  [0.71  log P(Ko/w)]  10.0061  MW, where kp is the permeability coefficient, P(Ko/w) is the partition coefficient in octanol compared to water, and MW is molecular weightg Toxicity Rat: Oral, male LD50: 13–15 mg/kg Skin, male LD50: 21 mg/kg Oral, female LD50: 3–3.6 mg/kg Skin, female LD50: 6.8 mg/kg a

Absorption of parathion has been established in humans, rat, mouse, frog, quail, and roach. Feldmann and Maibach (1974). c Marty (1976). d Maibach et al. (1974). e Shah et al. (1983). f Knaak et al. (1984). g Guy and Potts (1992). b

add analytic sensitivity. In human studies, plasma levels of compound are extremely low after topical application, often below assay detection level, so it is often necessary to use tracer methodology. The compound, usually labeled with 14C or tritium, is applied and the total amount of radioactivity excreted in urine or urine plus feces is

determined. The amount of radioactivity retained in the body or excreted by some route not assayed is corrected for by determining the amount of radioactivity excreted after parenteral administration. This final amount of radioactivity is then expressed as the percentage of applied dose that was absorbed (Feldmann and Maibach, 1974).

Chapter | 27  Agricultural Chemical Percutaneous Absorption and Decontamination

685

Distribution of systemic absorption

27.2.4  Stripping Method

Compound name: Parathion Dose: 4 µg/cm2 on whole body area (1.8 m2) continuous/infinite dose application Exposure: 24 hours

The stripping method determines the concentration of chemical in the stratum corneum during an application period and predicts the percutaneous absorption of that chemical. The chemical is applied to the skin of animals or humans, and at various skin application times the stratum corneum is removed by successive tape application and removal. The tape strippings are assayed for chemical content. Pharmacokinetic Cmax, Tmax, and area-under-the-curve parameters can be calculated for stratum corneum bioavai­lability (Nicoli et al., 2008; Rougier et al., 1986). Wu and Chiu (2007) used the tape stripping technique and quantitatively analyzed the tape-stripped samples utilizing an optical method (Fournier transform infrared spectroscopy) as an alternative to the conventional gas chromatography technique because the latter method is costly and time­consuming. The tape-stripped samples were grouped under four conditions: (1) the pesticide chlorpyrifos without the influence of the stratum corneum; (2) the sample of chlorpyrifos with the influence of stratum corneum (the forearm of one subject was the tape stripping site); (3) the mixture of chlorpyrifos and captan with the influence of the stratum corneum; and (4) the pesticide mixture without the influence of the stratum corneum. Calibration curves for each condition were performed. All spectra were imported into SAS (version 8, SAS Institute, Inc., Cary, NC), plus partial least squares algorithm and the principal component regression for supplementary spectral analysis. The results were evaluated by computing two indicators – percentage divergence (utili­ zing a mathematical equation) and precision (the coefficient of variation) of the samples tested (Tables 27.2 and 27.3).

Head and Neck 0.992 grams*

Trunk Front + Back 2.976 grams*

Arms and Hands (Left + Right) 0.124 grams

Genital 0.331 grams Legs and Feet (Left + Right) 0.496 grams

Total systemic absorption: 4.918 grams* Head, Neck and Arms = 1.116 grams* Estimated systemic LD50 of Parathion is 980 mg (human, 70 kg) *Indicates 50% lethality dose Figure 27.1  Simulated parathion human skin exposure to regions of the body. As early as 8 h following exposure, lethality is possible. At 24 h, lethality is possible if only certain body regions are exposed, such as the head and neck of a fieldworker.

The equation absorption is

used

to

determine

percutaneous

Total radioactivity after topical administration Absorption (%)   100 Total radioactivity after parenteral administration

27.2.3  Skin Flaps The methodology is to isolate surgically a portion of skin so that a singular blood supply is created to collect blood containing the chemical that has been absorbed through skin. The skin flap can be used to study percutaneous absorption in vivo or in vitro. The absorption of chemicals through skin and metabolism within the skin can be determined by assay of the perfusate (Wester and Maibach, 1997).

Table 27.2  Summary of the Quantification of Chlorpyrifos – Model Testing (with and without Stratum Corneum Influence)a % Divergence average (min, max)

Precision average (min, max)

R2

Low loading without SC

7.8 (2.8, 17.3)

5.7 (4.2, 7.4)

0.98

Low loading with SC

9.7 (4.7, 17.7)

6.4 (5.7, 7.8)

0.98

High loading without SC

5.1 (3.2, 8.6)

5.2 (1.3, 9.8)

0.98

High loading with SC

5.0 (3.7, 7.9)

4.2 (3.1, 5.5)

0.98

a

On average, high loading range of chlorpyrifos test samples had a better percentage divergence than low loading range, indicating that the presence of stratum corneum increases percentage divergence but not significantly. Adapted from Wu and Chiu (2007).

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Table 27.3  Summary of the Quantification of Chlorpyrifos and Captan – Model Testing (with and without Stratum Corneum Influence)a % Divergence average (min, max)

Precision average (min, max)

R2

  Chlorpyrifos

4.8 (1.3, 6.1)

5.8 (1.8, 8.4)

0.97

 �������� Captan

5.2 (5.2, 5.2)

6.3 (5.0, 7.0)

0.96

  Chlorpyrifos

9.6 (1.2, 16.3)

11.1 (1.6, 18.8)

0.90

 �������� Captan

6.1 (3.0, 10.5)

5.8 (3.3, 8.7)

0.98

Mixture without SC

Mixture with SC

a

Highest percentage divergence was associated with chlorpyrifos samples in the presence of stratum corneum, likewise indicating that the presence of stratum corneum increases percentage divergence but not significantly. Adapted from Wu and Chiu (2007).

The percentage divergence was mostly below 10%, except for several loading factors. The stratum corneum’s influence increased the percentage divergence but not significantly. R2 values for calibration curves and test samples were larger than 0.99 and 0.90, respectively. Details of the stripping method are described by Nicoli et al. (2008).

27.2.5  Biological Response Another in vivo method of estimating absorption is to use a biological or pharmacological response. Here, a biological assay is substituted for a chemical assay and absorption is estimated. An obvious disadvantage to the use of a biological response is that it is only good for compounds that will elicit an easily measurable response. An example of a biological response is the vasoconstrictor assay in which the blanching effect of one compound is compared to that of a known compound. This method is perhaps more qualitative than quantitative. The best known use of this method is in the comparison of corticosteroid products for dermatitis (Wester and Maibach, 1997).

27.2.6  In Vitro and In Vivo Methodologies In vitro percutaneous absorption is done with human and/or animal skin. The skin should be used as soon as possible. In vitro penetration gives mass results suitable for distinguishing drug formulation, especially in cases in which the drug will partition into reservoir fluid. Material balance in an in vitro study design adds to the overall data integrity. In vivo verification of skin absorption, preferably in humans, adds relevance to the in vitro data. The human skin sample can be kept viable if stored properly in the refrigerator (freezing kills skin viability) and used appropriately (Wester et al., 1984).

Table 27.4 gives the in vitro human skin and in vivo percutaneous absorption of several chemicals and vehicles. The in vitro absorption is divided into skin content and receptor fluid (either buffered saline or human plasma) accumulation. Receptor fluid accumulation does not necessarily agree with in vivo percutaneous absorption, perhaps because of minimal solubility in the receptor fluid. In some cases, skin content (see DDT) reflects in vivo absorption because the chemical was able to penetrate skin (and, lacking solubility, failed to partition into receptor fluid). Chemicals with high log P (octanol:water partition coefficient) minimally partition into receptor fluid (Wester and Maibach, 1997, 1999). Hostynek and Maibach (2005) provide data on advanced analytic methodology adding sensitivity to various methods. Boudry et al. (2008) determined and compared the percutaneous penetration and absorption of parathion using three experimental models – the human abdominal and pig ear skin in vitro models and the human skin onto a nude mouse (HuSki) in vivo model. The sample collection, treatment of samples, histological examination, and radioactivity measurements were described. The two in vitro skin models showed similar cumulative percutaneous penetration kinetic profiles in both acetone and ethanol as vehicles. The skin retention of the pig skin was two or three times lower (thus greater absorption) than the human abdominal skin, and it may be related to the pig skin’s anatomical structure. Pig model stratum corneum thickness is 8–13 m versus 10–17 m for human abdominal skin; likewise, the pig skin follicular diameter is 38–71 m versus 18 m for human abdominal skin. Dick and Scott (1992) suggested that if the ratio of absorption through the human skin is less than 3, as in this case and in other organophosphate studies (Tregear, 1966; Vallet et al., 2007), pig ear skin could be a suitable model for in vitro parathion percutaneous penetration studies. When ethanol was used as the vehicle, parathion absorption was four or five times higher in the HuSki model

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687

Table 27.4  In vitro versus In vivo Percutaneous Absorptiona Compound

Percentage dose In vitro

DDT

Benzo(a)pyrene

Chlordane

Pentachlorophenol

PCBs (1242)

PCBs (1254)

2,4 Dichlorophenoxy-

Log P a

Vehicle

Skin

Receptor fluid

In vivo

6.9

Acetone

18.1  13.4

0.08  0.02

18.9  9.4

Soil

  1.0  0.7

0.04  0.01

  3.3  0.5

Acetone

23.7  9.7

0.09  0.06

51.0  22.0

Soil

  1.4  0.9

0.01  0.06

13.2  3.4

Acetone

10.8  8.2

0.07  0.06

  6.0  2.8

Soil

  0.3  0.3

0.04  0.05

  4.2  1.8

Acetone

  3.7  1.7

  0.6  0.09

29.2  5.8

Soil

0.11  0.04

0.01  0.00

24.4  6.4

5.97

5.58

5.12

High

High

2.81

Acetone

21.4  8.5

TCB

18.0  8.3

Mineral oil

  6.4  6.3

  0.3  0.6

20.8  8.3

Soil

  1.6  1.1

0.04  0.05

14.1  1.0

Acetone

14.6  3.6

TCB

28.0  8.3

Mineral oil

10.0  16.5

  0.1  0.07

20.4  8.5

Soil

  2.8  2.8

0.04  0.05

13.8  2.7

Acetone

  8.6  2.1

acetic acid (2,4-D)

Soil

  1.6  0.2

0.02  0.01

159  4.7

Arsenic

Water

  1.0  1.0

  0.9  1.1

  2.0  1.2

Soil

  0.3  0.2

  0.4  0.5

  3.2  1.9

Water

  6.7  4.8

  0.4  0.2

Soil

0.09  0.03

0.03  0.02

Water

28.5  6.3

0.07  0.01

7.9  2.2

0.06  0.01

Cadmium

Mercury Soil a

6

Note that a log P of 6 means that 10 (1 million) molecules will partition into octanol for each molecule that will partition into water.

(31.7%) compared with previously published in vivo pig skin model investigations����������������������� – ���������������� 6.7 and 7.7% by ������ Qiao et al. (1993) and Carver and Riviere (1989), respectively���������������������� – validating ��������������� the usefulness of the HuSki model for mass balance studies. Using acetone as the vehicle, at 24 h post application, there was a close correlation between the parathion dose ��������� directly absorbed in the two human models [human in vitro (13.8%) and HuSki in vivo] and that of a published report by Maibach et al. (1971) in six human volunteers. The capability of the in vitro model to predict the human in vivo model could be due to the fact that in this study’s human in vitro model, the

rate-limiting step (i.e., the stratum corneum) maintains its barrier properties likewise in vivo (Boudry et al., 2008).

27.3  Regional variation in human and animal pesticide percutaneous absorption Feldmann and Maibach (1967) first systematically explored the potential for regional variation in percutaneous absorption in vivo in humans. The first absorption studies were

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Percentage dose absorbed

200

Forearm Palm Foot Abdomen Hand Fossa cubitalis Scalp Jaw Postauricular Forehead Ear canal Axilla Scrotum

100

0

Figure 27.2  Different parts of the body vary in percutaneous absorption. This is an important consideration in risk assessment. The scrotum has the greatest chemical absorption, followed by the head and neck.

done using the ventral forearm because this site is convenient to use. However, skin exposure to chemicals may exist over the entire body. They first showed regional variation with the absorption of parathion (Figure 27.2). The scrotum was the highest absorbing skin site (scrotal cancer in chimney sweeps was the key to identifying this fact). Skin absorption was lowest for the sole and highest around the head and face. Table 27.5 provides the effect of anatomical region on the percutaneous absorption of pesticides in humans (Maibach et al., 1971). There are two major points: first, regional variation was confirmed with the different chemicals. Second, those skin areas that would be exposed to agricultural chemicals, the head and face, were higher absorbing sites. Body areas most exposed to environmental contaminants are among the areas with the higher skin absorption. Table 27.6 demonstrates site variability for parathion skin absorption with time. Soap-and-water washes – even in the first few minutes after exposure������������������������� – are ������������������ not a perfect decontaminant. Site variation is apparent early in skin exposure (Wester and Maibach, 1985). Decontamination is discussed later. Guy and Maibach (1985), utilizing the hydrocortisone and pesticide data, constructed penetration indices for five anatomical sites (Table 27.7). The indices might be used with their total surface areas (Table 27.8) when estimating systemic availability relative to body exposure sites (Table 27.9) (Guy and Maibach, 1985). Van Rooy et al. (1993) applied coal tar ointment to various skin areas of volunteers and determined absorption of polycyclic aromatic hydrocarbons (PAHs) by surface disappearance of PAH and the excretion of urinary 1-OH pyrene. Using PAH disappearance, skin ranking (highest to lowest) was shoulder  forearm  forehead  groin   hand (palmar)  ankle. Using 1-OH pyrene excretion, skin

Table 27.5  Effect of Anatomical Region on in vivo Percutaneous Absorption in Humansa Anatomical region

Percentage of dose absorbed Hydrocortisone

Parathion

Malathion

Forearm

1.0

8.6

6.8

Palm

0.8

11.6

5.8

Foot, ball

0.2

13.5

6.8

Abdomen

1.3

18.5

9.4

—

21.0

12.5

Forehead

7.6

36.3

23.2

Axilla

3.1

64.0

28.7

12.2

33.9

69.9

—

28.4

4.4

32.1

—

46.6

36.2

101.6

Hand, dorsum

Jaw angle Fossa cubitalis Scalp Ear canal Scrotum a

Body areas most exposed to the environmental contaminant (head, face, and scrotum) are the areas with higher skin absorption.

ranking (highest to lowest) was neck  calf  forearm   trunk  hand. Table 27.9 compares their results with those of Guy and Maibach (1985). Wester et al. (1984) determined the percutaneous absorption of paraquat in humans. Absorption was the same for the leg (0.29  0.02%), hand (0.23  1%), and

Chapter | 27  Agricultural Chemical Percutaneous Absorption and Decontamination

Table 27.6  Site for Variation and Decontamination for Parathion dose absorbed (%)a Skin residence time before soap-andwater wash

Arm

1 min

2.8

Forehead

Palm

8.4

5 min

6.2

15 min

6.7

30 min

7.1

13.6

12.2

13.3

1 h

8.4

10.5

11.7

4 h

8.0

27.7

7.7

24 h

8.6

36.3

11.8

a Soap-and-water wash in the first few minutes after exposure is not a perfect decontaminant.

Table 27.7  Penetration Indices for Five Anatomical Sites Assessed Using Hydrocortisone Skin Penetration Data and Pesticide (Malathion and Parathion) Absorptiona Penetration index based on

Site

Hydrocortisone data

Pesticide data

40

12

Arms

1

1

Legs

0.5

1

Trunk

2.5

3

Head

5

4

Genitals

a The product of these indices and the body surface area might be used for systemic chemical absorption estimation.

689

forearms (0.29  0.1%). Here, the chemical nature of the low-absorbing paraquat overcame regional variation. Skin absorption in the rhesus monkey is considered to be relevant to that of humans. Table 27.10 shows the percutaneous absorption of testosterone (Wester et al., 1980), fenitrothion, aminocarb, and diethyltoluamide (DEET) (Moody and Franklin, 1987; Moody, Benoit et al., 1998) in the rhesus monkey compared with the rat. For the rhesus monkey, there is regional variation between forehead (scalp) and forearm. The ratio of forehead (scalp) to forearm for the rhesus monkey is similar to that for the human (Table 27.11). Therefore, the rhesus monkey may be a relevant animal model for human skin regional variation.

27.4  Percutaneous absorption from chemicals in clothing Chemicals in cloth cause cutaneous effects. For example, Hatch and Maibach (1986) reported that chemicals added to cloth in 10 finish categories (dye, wrinkle resistance, water repellency, soil release, etc.) caused irritation and allergic contact dermatitis, atopic dermatitis exacerbation, and urticarial and phototoxic skin responses. This is qualitative information that chemicals will transfer from cloth to skin in vivo in humans. All agricultural and chemigation (i.e., application of pesticides to irrigation) workers and the employers and employees of commercial pesticide establishments are directed to wear “personal protective equipment” clothing under the Worker Protection Standard by the U.S. EPA (http://www.epa.gov/pesticides/safety/workers/PART170. htm) when exposed to agricultural chemicals to reduce or eliminate pesticide contamination (Driver et al., 2007). The clothing required depends on the extent of pesticide exposure. In general, fabric coveralls over a long-sleeved shirt and long pants are recommended for most toxic products; chemical-resistant footwear and gloves are added

Table 27.8  Body Surface Area Distributed over Five Anatomical Regions for Adults and Neonates Anatomical region

Adult Body area (%)a

Genital

Neonate Area (cm2)

Body area (%)

Area (cm2)

1

180

1

19

Arms

18

3,240

19

365

Legs

36

6,480

30

576

Trunk

36

6,480

31

595

Head

9

1,620

19

Total a

18,000

Note the “rule of 9” when trying to remember the human body surface area.

365 1920

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690

Table 27.9  Absorption Indices of Hydrocortisone and Pesticide (Parathion/Malathion) Calculated by Guy and Maibach (1985) Compared with Absorption with Pyrene and PAH for Different Sites by Van Rooy et al. (1993) Absorption index Anatomical site

Hydrocortisonea

Pesticidesb

Pyrenec

PAHd

Genitals

40

12

Arms

1

1

1

1

Hands

1

1

0.8

0.5

—

—

Legs, ankle

0.5

1

1.2

Trunk, shoulder

2.5

3

1.1

—, 2.0

0.8, 0.5

Head, neck

5

4

—, 1.3

1.0

a

Based on hydrocortisone penetration data (Feldmann and Maibach, 1967). Based on parathion and malathion absorption data (Maibach et al., 1971). Based on the excreted amount of 1-OH-pyrene in urine after coal tar ointment application (Van Rooy et al., 1993). d Based on the PAH absorption rate constant (Ka) after coal tar ointment application (Van Rooy et al., 1993). b c

Table 27.10  Percutaneous Absorption of Fenitrothion, Aminocarb, DEET, and Testosterone in Rhesus Monkey and Rat Chemical

Fenitrothion

Applied dose absorbed

Species

Rhesus

Forehead

Forearm

49

21

Rat Aminocarb

Rhesus

37

Rat Testosterone

Rhesus

88 20.4a

8.8 47.4a

Rat DEET

Rhesus Rat

33

Chemical

Percutaneous absorption ratio

Species

Back

84 74

Table 27.11  Percutaneous Absorption Ratio for Scalp and Forehead to Forearm in Humans and Rhesus Monkeysa

14 36

a

Scalp.

Scalp/ forehead

Forehead/ forearm

3.5

6.0

Hydrocortisone

Human

Benzoic acid

Human

Parathion

Human

Malathion

Human

Testosterone

Rhesus

Fenitrothion

Rhesus

2.3

Aminocarb

Rhesus

2.0

DEET

Rhesus

2.4

2.9 3.7

4.2 3.4

2.3

a

The ratio of the forehead (scalp) of rhesus monkey is comparable with that of humans, indicating that the rhesus monkey may be a relevant animal model for human skin regional variation studies.

if a chemical has a high dermal toxicity or skin irritation potential; and the single-layer clothing (long-sleeved shirt and long pants) is recommended for low-risk pesticide exposure (Driver et al., 2007). Driver et al. (2007) performed a quantitative characterization of the penetration of pesticide chemical residues in various types and configuration of clothing using the U.S. EPA’s Pesticide Handlers Exposure Database (PHED). The main objective of this cohort study was to develop pesticide clothing penetration (or, conversely, protection) factors for singlelayer clothing based on the dermal exposure monitoring

data [obtained from passive dosimetry (dosimeter attached to the inside of the clothing as opposed to outer dosimetry, where the meter is attached outside the clothing) of patch dosimeters values]. For estimating the potential skin exposure from the passive dosimeter data, the penetration factor was represented as the fraction of pesticide that crosses the barrier of single-layer clothing and is available for skin contact. Percentage clothing penetration (%CP) was investigated as a function of (1) body part (patch vs. whole body), (2) application method, (3) formulation type used in mixing/

Chapter | 27  Agricultural Chemical Percutaneous Absorption and Decontamination

691

Table 27.12  Mean Percentage Clothing Penetration for Single-Layer Clothing by the Type of Application Methoda

Table 27.13  In vivo Percutaneous Absorption of Glyphosate and Malathion from Cloth through Human Skina

Application method (code no.)

Chemical

Donor condition

Treatment

% of dose absorbed

Glyphosate

1% solution (water)

None

1.42  0.25

1% solution on cloth

0 h

0.74  0.26

Not specified (0)

Mean

n

95% Confidence interval of mean

7

14.38

3.98–24.77

Airblast (1)

403

8.53

7.40–9.65

Groundboom tractor (2)

178

11.00

8.92–13.07

1% solution on cloth

24 h

0.08  0.01

Groundboom truck (3)

22

18.26

9.62–26.89

1% solution on cloth

48 h

0.08  0.01

180

14.62

12.68–16.56

Add water

0.36  0.07

25

20.41

12.04–28.79

1% solution on cloth

187

10.99

9.16–12.82

1% solution (water/ethanol)

None

8.77  1.43

Paint brush (8)

75

10.49

6.63–14.34

1% solution on cloth

0 h

3.92  0.49

Backpack sprayer (9)

50

8.69

5.36–12.03

0.62  0.11

4.14

2.31–5.97

1% solution on cloth

24 h

105

Rights-of-way sprayer (11)

40

24.97

19.27–30.68

1% solution on cloth

48 h

0.60  0.14

High-pressure hand wash (greenhouse and ornamentals) (13)

43

14.20

10.40–17.99

1% solution on cloth

Add water

7.34  0.61

Aerosol can (4) Aerial–fixed wing (5) Low-pressure hand wand (7)

Airless sprayer (10)

a

Termiticide injection (16)

106

10.73

7.41–14.06

Solid broadcast spreader (belly grinder) (18)

139

10.77

8.83–12.72

6

8.71

1.10–16.32

Hand dispersion, granular bait (22)

Malathion

a

The application method with the highest mean percentage clothing penetration is the rights-of-way sprayer, whereas the lowest is the airless sprayer. Adapted from Driver et al. (2007).

loading (solid and dry), and (4) job classification. Although there were no significant differences by body part or by formulation for %CP, there were larger differences in values for the application methods (Table 27.12). Grand mean singlelayer clothing penetration values for patch (n  2029) and whole-body (n  100) dosimeter samples from PHED were 12.12% (SE, 0.33; SD, 15.02) and 8.21% (SE, 1.01; SD, 10.14), respectively. Regression analysis for all inner and outer dosimetry samples (the log-transformed of the paired values) reflected the hypothesis that the amount of pesticide in single-layer clothing appears to increase with decreasing outer dosimeter loading or challenge.

Both glyphosate and malathion in solution (treatment  none) are absorbed through human skin. Glyphosate and malathion on cotton cloth show absorption in skin, depending on the time the chemical was added to the cloth (treatment  0, 24, and 48 h). When the cloth was wetted (treatment  add water/ethanol), the transfer of glyphosate and malathion from cloth to human skin was increased. This suggests that sweating, skin oil, or even rain may facilitate transfer of chemicals from cloth to skin.

In other studies (Wester et al., 1996), in vitro percutaneous absorption of glyphosate and malathion through human skin was decreased when they were added to cloth (and the cloth then placed on skin), and this absorption decreased further after 48 h (Table 27.13). It is assumed that with time, the chemical will sequester into deep empty spaces of the fabric, or some type of bonding will be established between chemical and fabric. When water was added to glyphosate–cloth and water/ethanol to malathion–cloth, the percutaneous absorption increased (malathion to levels from solution). This perhaps reflects clinical situations in which dermatitis occurs most frequently in human sweating areas (axilla and crotch). The clothing must not be a collection system for pesticides, and it cannot be assumed that laundering will remove the agents. Improper laundering of pest control operator overalls may allow pesticide dermal absorption during subsequent use (Stone and Stahr, 1989). Chlorine bleach pretreatment offers the advantage of oxidative degradation of

692

organophosphates – for example, chlorpyrifos degraded into 3,5,6-trichloro-2-pyridinol (Perkins et al., 1996). Laundering protocols for the organophosphate chlorpyrifos residue removal were investigated utilizing three chlorpyrifos-containing pesticides: Dursban PC Termiticide and Insecticide (PCT) (emulsifiable concentrate), Dursban Micro-Lo Termiticide (emulsifiable concentrate), and Empire Insecticide (capsule suspension formulation) (Dow AgroScience, Inc., Indianapolis, IN) – with fenthion (Baytex, Bayer CropScience, Victoria, Australia) and permethrin (Perigen, Bayer CropScience) included for comparison purposes (Table 27.14) (Fitzgerald and Manley-Harris, 2005). The type of fabric overalls, amount of pesticides applied, type of washing machine, and extraction of residual pesticide laundered fabrics were clearly explained. The recommended amount on the label of each detergent/chemical was followed. There was significantly greater chlorpyrifos retention in the polyester fabric versus 100% cotton material in preliminary trials using Dursban PCT (trials 1–9); thus, 100% cotton material was utilized in the succeeding trials. Trials 1–4 and 11 suggest that hot or cold washes and the quality of detergent did not significantly influence recovered chlorpyrifos. Significantly greater chlorpyrifos was recovered with the Dursban Micro-Lo preparation than with Dursban PCT (trials 2, 13, and 14): although both are emulsifiable concentrates, the former has smaller droplets, providing a higher surface proportion of surfactant. Overnight soak with sodium percarbonate (NapiSan, Reckitt Benckiser, West Ryde, NSW, Australia) offered significant protection (trials 15 and 16) because low levels of chlorpyrifos residual were obtained, suggesting that chlorpyrifos degradation is also hastened in alkaline conditions. The polymer coating of the chlorpyrifos granules may have prevented Empire’s adherence to the fabric (trials 17 and 18). There was a significant increase in chlorpyrifos residues in the trials with fenthion and permethrin (trials 19–22). The following laundering protocols regarding chlorpyrifos-contaminated overalls were recommended: (1) 100% cotton material should be laundered after single use by first soaking for 24 h in NapiSan or any product with an equivalent amount of sodium percarbonate (using the recommended quantity printed on its label), at least initially in hot water, and then soak water should be discarded; (2) the material should be washed for 15 min or longer with appropriate water level and amount of laundry detergent (at the recommended label use rate); and (3) despite the fact that the polymer-coated chlorpyrifos granules (Empire) did not degrade when washed with sodium percarbonate, it is still recommended to presoak contaminated overalls to deal with possible chlorpyrifos release in cases of damaged polymer coating. This protocol may be extended to fenthion formulations, but it is not recommended for permethrin. Boman et al. (2004) discuss protective gloves and clothing.

Hayes’ Handbook of Pesticide Toxicology

27.5  Models for agricultural chemical assessments and predictions 27.5.1  The Cluster Analysis Method The assumption based on the absorption from a singlesolvent system may be inappropriate for risk assessment when dealing with chemical mixtures (Van der Merwe and Riviere, 2006). Cluster analysis is used to evaluate a large number of treatments and is useful when similar numerical data points prevent intuitive identification of data structure. It also identifies clusters in a data set that are distinct from each other based on mathematical indices of similarity and dissimilarity (Van der Merwe and Riviere, 2006). Knowing the relationships between treatments that have similar effects on absorption could form bases for hypothesis generation (Van der Merwe and Riviere, 2006). Van der Merwe and Riviere (2006) utilized the cluster analysis method to establish that the determining influence of the solvent polarity on the partitioning data structure supported the hypothesis that solvent polarity drives the partitioning of nonpolar solutes – utilizing 10 chemicals (phenol, p-nitrophenol, pentachlorophenol, methyl parathion, parathion, chlorpyrifos, fenthion, simazine, atrazine, and propazine) in 24 chemical mixtures – in female weanling porcine skins. Kmeans and hierarchical cluster analyses (using MATLAB version 6.5.0.180913a, release 13; Mathworks, Inc., Natick, MA) were obtained from (1) the stratum corneum/solvent partitioning data, (2) permeability data, and (3) polarity indices of the 24 solvent systems. Partitioning coefficient (log P) was determined by the log of the normalized radioactivity of the content in the vehicle mixture (Cvehicle mixture) and in the stratum corneum (Cstratum corneum). On the other hand, permeability was estimated by dividing the slope of the steady-state portion of the cumulative mass absorbed/time curve by the concentration in the donor solvent. Lastly, the polarity index of each solvent system was created by summing the products of the log P values of each component and their proportional contributions to the total mass of the solvent system (mass of component/mass of solvent system). A clustering structure based on solvent polarity became apparent from the stratum corneum/solvent partitioning data (Tables 27.15 and 27.16). Hierarchical cluster 4 may be described as substantially polar (average polarity index of 0.001), cluster 3 as mildly polar (average polarity index of 0.235), cluster 2 as substantially nonpolar (average polarity index of 0.727), and cluster 1 as mildly nonpolar (average polarity index of 0.369). Molecules with similar polarity have relatively higher intermolecular attraction than molecules with dissimilar polarity. Thus, under controlled conditions, solute molecules in a solvent system of dissimilar polarity exist in a state of higher potential energy compared with solute molecules in a solvent system of similar polarity. Due to the effects of intermolecular forces on enthalpy (i.e., heat content), equilibrium of energy was reached when partitioning into the

Chapter | 27  Agricultural Chemical Percutaneous Absorption and Decontamination

693

Table 27.14  Percentage Residual Chlorpyrifos from Washing Protocolsa Trial

Procedure

% Residual chlorpyrifos Polyester/ cotton

100% Cotton

Preliminary trial with Dursban PCTb Cold washc with Persild concentrate

  1

e

43.4

9.0

  2

Hot wash with Persil concentrate

34.7

11.6

  3

Cold wash with liquid Drivef concentrate

45.4

11.6

  4

Cold wash with liquid Drive washing powder

27.8

11.7

  5

Reduced cold wash with Persil concentrate

56.4

16.8

  6

Bleach soak (1.5 h) prior to cold wash with Persil concentrate

33.6

8.8

  7

Bleach soak (4 h) prior to cold wash with Persil concentrate

10.7

3.7

  8

Sunlight exposure prior to cold wash with Persil concentrate

68.9

16.5

  9

Repeated applications of Dursban PTC prior to cold wash with Persil concentrate

35.1

13.1

Further trials with Dursban PCT   10

Cold wash with no detergent

22.2

  11

Cold wash with budget (generic) washing powder

  12

Water soak (4 h) prior to cold wash with Persil concentrate

10.1

  13

Cold wash with Persil concentrate

14.6

  14

Cold wash with Persil concentrate

17.5

  15

Overnight soak in NapiSanh prior to cold wash with Persil concentrate

1.1

Overnight soak in Napisan prior to cold wash with Persil concentrate

1.3

  17

Cold wash with Persil concentrate

5.0

  18

Cold wash with Persil concentrate

4.7

  19

Cold wash with Persil concentrate

28.3

  20

Cold wash with Persil concentrate

31.2

  21

Cold wash with Persil concentrate

14.5

  22

Cold wash with Persil concentrate

14.5

9.4

Trials with Dursban Micro-Log

  16 i

Trials with Empire

Trials with fenthion

Trials with permethrin

a Overnight soak in NapiSan (sodium percarbonate) prior to cold wash with Persil concentrate (sodium percarbonate) protocol yielded the least percentage residual chlorpyrifos. b Dursban PC Termiticide and Insecticide, Dow AgroScience, Inc., c Wash cycle (15 min/1415°C), rinse cycle (15 min/1415°C), spin cycle (15 min). d Persil: active ingredient, sodium percarbonate; Johnson Diversey, Sturtevant, WI. e Same wash cycle at 60°C. f DRIVE: active ingredient, enzyme technology; UNILEVER, Epping, Australia. g Dursban Micro-Lo Termiticide, Dow AgroScience, Inc., Indianapolis, IN. h NapiSan: active ingredient, sodium percarbonate; Reckitt Benckiser, West Ryde, NSW, Australia. i Empire Insecticide, Dow AgroScience, Inc., Indianapolis, IN. Adapted from Fitzgerald and Manley-Harris (2005).

Hayes’ Handbook of Pesticide Toxicology

694

Table 27.15  Clustering of K-Means Based on Stratum Corneum/Solvent Partitioninga Solvent

Cluster

Water

4

Water  MNA

4

Water  PG

3

Water  PG  MNA

3

Water  PG  MNA  SLS

2

Water  PG  SLS

2

PG  MNA  SLS

2

PG  SLS

2

Ethanol  PG  MNA  SLS

2

Ethanol  PG  SLS

2

Ethanol  PG  MNA

2

Ethanol  PG

2

Ethanol  water  MNA  SLS

2

Ethanol  water  SLS

2

Ethanol  MNA  SLS

2

Ethanol  SLS

2

Ethanol  MNA

2

Ethanol

2

PG  MNA

1

PG

1

Water  MNA  SLS

1

Water  SLS

1

Ethanol  water  MNA

1

Ethanol  water

1

a

The clustering structure of the 24 solvents became apparent from the stratum corneum/solvent partitioning. MNA, methyl nicotinic acid; PG, propylene glycol; SLS, sodium lauryl sulfate. Adapted from Van der Merwe and Riviere (2006).

nonpolar environment of the stratum corneum lipids of relatively nonpolar compounds was higher from polar than from nonpolar solvents.

27.5.2  Determinants of Dermal Exposure Ranking Method (Derm): A Method for   Pesticide Exposure Assessment for Developing Countries Farmers in most developing countries have greater risk of pesticide exposure because, usually, improper pesticide

Table 27.16  Average Polarity Indices Based on Stratum Corneum Partitioning for Hierarchical and K-Means Clusters Average polarity indexa

Range

  4

0.001

0.000–0.003

  3

0.235

0.234–0.237

  2

0.764

0.480–1.070

0.325

0.271–0.473

  4

0.001

0.000–0.003

  3

0.235

0.234–0.237

  2

0.727

0.463–1.072

  1

0.369

0.271–0.462

K-means clusters

  1 Hierarchical clusters

b

a

Polarity index was calculated and used to quantify relative polarity of various solvent systems. b Hierarchical cluster 4 may be described as substantially polar, cluster 3 as mildly polar, cluster 2 as substantially nonpolar, and cluster 1 as mildly nonpolar. Adapted from Van der Merwe and Riviere (2006).

control techniques [including the use of backpack pesticide sprayers, the lack of proper protective clothing and equipment (e.g., refilling backpack sprayers using bare hands), exposure of pesticide from a leaking container while spraying], and inaccurate dermal pesticide exposure assessment are the only ones economically permissible (Blanco et al., 2008). The determinants of dermal exposure ranking method (DERM) was developed by Blanco et al. (2008) for use in developing countries. DERM is a model in which determinants of dermal exposure are assessed on the basis of two factors: the type of transport process and body surface area (BSA) affected. The transport process (T) is further classified into three categories – transfer, deposition, and emission – each of which is defined and given a corresponding score (Table 27.17). On the other hand, the BSA (A) is based on the percentage BSA estimates of Lund and Browder (1944) for burned patients; each percentage BSA estimate is also given a corresponding score (Table 27.18). The clothing of farmers was considered as a “protection factor” because even normal clothing offers protection from exogenous chemicals. The degree of the clothing protection factor (C) was defined as the complement of the reduction in the exposure level (1-exposure reduction), where the maximum exposure reduction considered was when the best clothing was worn (i.e., long-sleeved shirt and long pants), equivalent to 50% (expressed as 0.50) exposure reduction (see Table 27.18). DERM was tested in a small population of Nicaraguan subsistence farmers, and

Chapter | 27  Agricultural Chemical Percutaneous Absorption and Decontamination

Table 27.17  Scores for Category Factors: Transport (T) and Body Surface Area (A)

Table 27.18  Clothing Exposure Reduction in Nicaraguan Subsistence Farmersa

Factor

Category

Subcategory (example)

Scorea

Clothing

Transport (T)

Transfer

Touching a contaminated surface or any contact with contaminated clothing

1

Shirt

Touching a recently spilled/ overflowed/ splashed surface

3

Spraying against the direction of the wind or walking into a cloud of pesticide spray

4

Bare hands dipped inside pesticide tank and/or when fixing the nozzle

5

Deposition

Emission

Body surface area (A) (%)b

695

Body surface area covered (%)

Exposure reduction

 Worn-out/torn/ overused

0

0

  Short-sleeved

33

0.15

  Long-sleeved

42

0.20

 Worn-out/torn/ overused

0

0

  Short

25

0.10

  Long

39

0.20

  Yes

7

0.10

  No

0

0

Pants

Shoes

a

Clothing protection factor (C) was defined as the complement of the reduction in the exposure level (1 – exposure reduction). Adapted from Blanco et al. (2008).

Not applicable

1

81–100

5

61–80

4

41–60

3

21–40

2

0–20

1

Not applicable

1

a A score of 5 indicates greater chemical exposure, whereas a score of 1 indicates a “nonapplicable” exposure. b Based on Lund and Browder’s (1944) burnt patient estimates. Adapted from Blanco et al. (2008).

the resulting determinants were ascertained. DERM was compared with two recently developed semiquantitative methods – the total visual score (TVS) and the contaminated body area (CBA). DERM was in agreement with both TVS (r  0.69; p  0.000) and CBA (r  0.67; p  0.000).

27.5.3  Dermal Assessment Estimate (Dream) Method Van Wendel de Joode et al. (2005) developed a semiquantitative technique, the dermal assessment and estimate

method which estimates the amount of chemical in the clothing layer and the skin and provides insight into the distribution of dermal exposure in the body. It requires an inventory of the occupational setting and is performed in two parts: (1) a multiple-choice questionnaire (regarding the dermal route/dermal covering and protection/exposure duration/physical and chemical properties of the chemical) and (2) the mathematical evaluation of the answers. The outcome is a numerical value that is classified according to the following DREAM risk categories: 0, no risk; 1–10, very low risk; 11–30, low risk; 31–100, moderate risk; 101–300, high risk; 301–1000, very high risk; 1000, extremely high risk (van Wendel de Joode et al., 2005).

27.5.4  Estimation and Assessment Exposure (Ease) Model The UK Health and Safety Executive and the Health and Safety Laboratory developed the estimation and assessment exposure model (EASE), which is a computer-based software program for inhalational and dermal exposure utilizing data from the HSE’s National Exposure Database, with the inhalational exposure assessment being more developed and comprehensive than the dermal exposure assessment (Tickner et al., 2005). It was designed to provide outputs of broad estimates utilizing a series of multiplechoice questions, and the estimates are solely based on

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the answers (Boogaard, 2008). The dermal exposure is evaluated as the potential exposure rate to the hands and the forearms (a total skin area of approximately 2000 m2) (Boogaard, 2008).

27.5.5  Risk Assessment of Occupational Dermal Exposure (Riskofderm) Funded by the European Commission (RISKOFDERM, QLKA4-CT-1999-01107), 15 European scientists were tasked to develop (1) a validated predictive model for estimating dermal exposure for use in generic risk assessment for single chemicals and (2) a practical dermal exposure risk assessment and management toolkit for use by small and medium-sized enterprises and others in actual workplace situations. They developed the Risk Assessment of Occupational Dermal Exposure (RISKOFDERM) (van Hemmen et al., 2003). RISKOFDERM is a predictive model for estimating dermal exposure for use in risk assessment of a single chemical as well as a dermal exposure risk management tool for workplaces (Boogaard, 2008).

27.5.6  Biosensors The conventional methods for detecting organophosphates (e.g., gas chromatography, high-performance liquid chromatography, and thin-layer chromatography) bear the following disadvantages: (1) the systems are highly complex, (2) the procedures are time-consuming, (3) highly trained personnel are required, and (4) the methods are too costly. On the other hand, the use of biosensors as an alternative technique to conventional methods is cost-effective, the technique is simple to perform, and it permits in situ detection of pesticides. The acetylcholinesterase inhibition-based biosensors are widely used for detecting organophosphate compounds and are especially sensitive to chlorpyrifos and chlorfenvinfos. Istamboulie et al. (2009) added the biocatalyst phosphotriesterase to the highly sensitive recombinant Drosophila melanogaster acetylcholinesterase (B394) for the selective detection of chlorpyrifos and chlorfenvinfos insecticides through amperometry – measurement based on the use of electric current. The inhibition percentage was correlated with the concentration of the insecticide, and the limit of detection was calculated as the concentration of the insecticide generating a 10% decrease in the biosensor response. The combination of high sensitivity of B349 acetylcholinesterase with phosphotriesterase to hydrolyze organophosphate compounds, and as such to use as a biosensor selective to chlorpyrifos and chlorfenvinfos, was demonstrated in this study because both chlorpyrifos and chlorfenvinfos chemicals have high affinity to phosphotriesterase. Although chlorfenvinfos has a higher affinity to phosphotriesterase than chlorpyrifos, chlorfenvinfos hydrolysis was shown to be 10,000-fold slower than chlorpyrifos.

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The addition of “blank” assays without phosphotriesterase substantiated that chlorfenvinfos was not hydrolyzed. Thus, the high affinity and low degradation velocity indicate that chlorfenvinfos may act as a competitive inhibitor of phosphotriesterase hydrolysis of other organophosphate compounds, thereby preventing the efficient hydrolysis of other organophosphates. To ensure that there was actual inhibition of the chemical and that it was not due to leakage, the stability of the biosensor was determined via successively measuring the response of B349 electrode to 1 mM acetylcholine at 100 mV versus Ag/AgCl.

27.6  Biomonitoring: tool for human health risk characterization Boogaard (2008) enumerated the elements that make up the human health risks characterization: (1) hazard identification (knowing the qualitative nature of the contaminant and potential adverse effect), (2) dose–response analysis (the relationship between dose and the incidence of its adverse effect), and (3) exposure risk assessment (dose of contaminant that will be acquired by the individual). In practice, the hazard and dose–response assessments are derived only from experimental animals because human studies are scarce. On the other hand, exposure risk assessment relies on the net deposition of a substance into the skin, through the extraction of the substance accordingly from (1) the surrogate skin method (e.g., gloves), (2) tape stripping, (3) visualization (e.g., use of fluorescent markers), and (4) washing the appropriate solvent. In occupational settings, most of the inhalational exposure risks have been recognized. The development of dermal exposure risk assessments (e.g., DREAM, EASE, and RISKOFDERM) has gained interest. These are classified as “external metrics” because they mainly deal with measuring the percutaneous penetration of chemicals and, hence, do not reflect actual chemical exposure. On the other hand, “internal metrics” refers to the percutaneous absorption of the chemical, reflecting a better view of the actual chemical exposure. Biomonitoring integrates dermal absorption and percutaneous penetration of a certain chemical through biological and biochemical effect monitoring, taking into account the metabolism including differences in susceptibility, providing a more direct link to health effects. It also integrates the different pathways of exposure (inhalation, oral, and dermal) – an important limitation of the external metrics (Boogaard, 2008).

27.7  Skin decontamination Decontamination of a chemical from the skin is commonly done by washing with water only or with soap and water because it has been assumed that washing will remove

Chapter | 27  Agricultural Chemical Percutaneous Absorption and Decontamination

the chemical, as demonstrated by Wester et al.’s (1991) glyphosate skin decontamination study utilizing water only and water  50% soap (Ivory liquid soap; Procter and Gamble Company, Cincinnati, OH) on the abdominal skin of rhesus monkeys in vivo. The abdominal skin was marked and dosed with 7 l/cm2 solution containing 0.4 g of [14C]glyphosate and then washed with water only or water  50% soap; after 5 min, the site was rinsed twice with water only. In the grid method (i.e., the entire abdominal skin was marked with 1 cm2), each 1-cm2 area was dosed with the same amount of [14C]glyphosate. At designated times (0, 0.5, 3, 6, and 24 h), the sites were washed with the decontaminating liquids. With water-only wash, 83.6  3.3% of the applied dose was removed in the single-site application, whereas 76.9  3.6% was significantly ( p  0.03) removed in the grid system. At 24 h, the amount of [14C]glyphosate recovered did not differ significantly in both washing solutions. Wester et al. (1992) performed another decontamination study utilizing water only and water and soap in washing the herbicide alachlor. Figure 27.3 illustrates skin decontamination of alachlor with soap and water or with water only during a 24-h dosing period using grid methodology. Note that the amount recovered decreases over time, which happens because this is an in vivo system and percutaneous absorption occurs, decreasing the amount of chemical on the skin surface. There also may be loss due to skin desquamation. A second observation is that alachlor is more readily removed with soap-and-water wash than with water only. The reason is that alachlor is lipid soluble and needs the surfactant system for more successful decontamination (Wester et al., 1991, 1992). In the preceding ­ illustration,

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the decrease in alachlor wash recovery over time was thought to be due to ongoing absorption and loss due to skin desquamation. These factors probably contribute but are probably not the main reason, which is soap-and-water wash effectiveness. In the home and workplace, decontamination of a chemical from skin is traditionally done with a soap-and-water wash, although some workplaces may have emergency showers. It has been assumed that these procedures are effective, yet workplace illness and even death occur from chemical contamination. Zhai et al. (2008) compared tap water with isotonic saline and hypertonic saline solutions for the removal of glyphosate in human cadaver skin. Each skin was dosed with approximately 375 g of [14C]glyphosate with varying exposure times (1, 3, and 30 min); the skins were then washed three times with 4 ml of each of the decontaminating solutions. Two tape strippings were performed post washing. The wash solutions, tape disks, receptor fluid, skin strippings, and the remainder of the skin were counted with a liquid scintillation analyzer to determine glyphosate mass. The total glyphosate mass balance for each group ranged from 94.8 to 102.4%. There were no statistical differences ( p  0.05) among the groups (Zhai et al., 2008). Water, or soap and water, may not be the most effective means of skin decontamination, particularly for lipid-soluble materials. A study was undertaken to help determine whether there are more effective means of removing methylene bisphenyl isocyanate from skin (Wester et al., 1999). MDI is an industrial chemical for which skin ­decontamination using traditional soap and water and nontraditional polypropylene 100

100

80 Percentage dose

Percentage dose

80

60

40

40

20

0

60

20

0

1

3

6

24

Time (hours) Soap and water

Water only

Figure 27.3  Washing of alachlor: soap and water versus water only. Alachlor is a lipophilic chemical that is better removed by soap and water than by water only.

0

2

4 6 Time (hours)

8

Water-only 5% soap

Polypropylene Dtam

50% soap

Corn oil

10

Figure 27.4  Mean percentage of applied dose MDI removed with designated decontamination procedure at designated time period. Water and the combination of soap and water are the least effective, especially at 4 and 8 h.

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Skin wash-in effect

Hydration effect

Surfactant effect

Friction effect

Acid/base effect

Artefact effect

Figure 27.5  Five most likely mechanisms to explain the wash-in effect. These five consequential effects of skin decontamination may enhance percutaneous chemical penetration and absorption (adapted from Moody and Maibach, 2006).

glycol, a polyglycol-based cleanser (DTAM), and corn oil was assayed in vivo on the rhesus monkey over 8 h (Figure 27.4). Water, alone or with soap (5 and 50% soap), was partially effective in the first hour after exposure, removing 51–69% of the applied dose. However, decontamination decreased to 40–52% at 4 h and 29–46% at 8 h. Thus, the majority of MDI was not removed by the traditional soap-and-water wash; skin tape stripping after washing confirmed that MDI remained on the skin. In contrast, polypropylene glycol, DTAM, and corn oil all removed 68–86% of the MDI in the first hour, 74–79% at 4 h, and 72–86% at 8 h. Statistically, polypropylene glycol, DTAM, and corn oil were all better ( p  0.05) than soap and water at 4 and 8 h after dose application. These results indicate that a traditional soap-and-water wash and the emergency water shower are relatively ineffective at removing MDI from skin. More effective decontamination procedures as discussed here are available. These procedures are consistent with the partial miscibility of MDI in corn oil and polyglycols (Wester et al., 1999). Thus, if there is skin contamination with an aqueous chemical and the skin is washed with soap and water, it cannot be assumed that the chemical has been removed from the skin. Evidence suggests that often the skin and the body are unknowingly subjected to enhanced penetration and systemic absorption or toxicity because the decontamination procedure does not work or may actually enhance absorption – a phenomenon called the “wash-in” (W-I) effect (Figure 27.5) (Moody and Maibach, 2006). Moody and Maibach reviewed articles that relate to skin absorption and decontamination, and they suggested the following mechanisms of the W-I effect: (1) the effects of skin hydration [e.g., the decontaminating surfactants (sodium dodecyl sulfate and benzathine chloride) and saline solutions (hypotonic, isotonic, and hypertonic) used in decontaminating diethyl malonate (Loke et al., 1999)]; (2) the surfactant effects on skin barrier integrity [i.e., resulting from skin irritation (e.g., the ionic surfactant sodium lauryl sulfate), membrane fluidization, or delipidation]; (3) friction (e.g., the use of a Q-tip dipped in the decontaminating solution to wash the stratum corneum); (4) acid–base reactions; and (5) artefact effects. The latter two have yet to be investigated further.

Conclusion Aqueous chemical use can achieve its chemically intended goals, but more knowledge of human risk assessment is required. Understanding percutaneous absorption as a major route of pesticides entering the body is an integral part of the risk assessment process. Data on humans can be obtained safely using trace measurement methodology and with low-risk doses coupled with high-tech analytical methodology. Although the data from animal and computer models are simpler to use, the method of choice is biomonitoring (Boogaard, 2008). For developing countries, DERM may be used for dermal exposure assessment (Blanco et al., 2008). Safety is debatable if the models are not validated to humans because the resulting risk asses­ sment may also be wrong. Protective clothing serves to prevent agricultural workers from unnecessary pesticide exposure. A protective layer or barrier between the worker and chemical contamination (1) prevents chemical penetration as a bulk flow through a porous material, (2) uses nonporous material with low permeation, and (3) absorbs or retains chemical in the fabric (Obendorf et al., 2003). Equally important is the proper laundering of clothing because improper laundering may allow pesticide absorption in subsequent use (Stone and Stahr, 1989). This chapter summarized what is known. As more is learned about human and animal skin, we should be able to more efficiently protect humans from potential adverse effects of aqueous chemical exposure.

References Blanco, L. E., Aragon, A. et al. (2008). The determinants of dermal exposure ranking method (DERM): a pesticide exposure assessment approach for developing countries. Ann. Occup. Hyg. 52(6), 535–544. Boman, A., Maibach, H. I. et al. (2004). “Protective Gloves for Occupational Use”. CRC Press, Boca Raton, FL. Boogaard, P. J. (2008). Biomonitoring as a tool in the human health risk characterization of dermal exposure. Hum. Exp. Toxicol. 27(4), 297–305. Boudry, I., Blanck, O. et al. (2008). Percutaneous penetration and absorption of parathion using human and pig skin models in vitro and

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Agricultural Chemical Percutaneous Absorption and Decontamination

human skin grafted onto nude mouse skin model in vivo. J. Appl. Toxicol. 28(5), 645–657. Carver, M. P., and Riviere, J. E. (1989). Percutaneous absorption and excretion of xenobiotics after topical and intravenous administration to pigs. Fundam. Appl. Toxicol. 13(4), 714–722. Dick, I. P., and Scott, R. C. (1992). Pig ear skin as an in-vitro model for human skin permeability. J. Pharm. Pharmacol. 44(8), 640–645. Driver, J., Ross, J. et al. (2007). Derivation of single layer clothing penetration factors from the pesticide handlers exposure database. Regul. Toxicol. Pharmacol. 49(2), 125–137. Farahmand, S., and Maibach, H. I. (2008). Transdermal drug pharmacokinetics in man: Interindividual variability and partial prediction. Int. J. Pharmaceutics 367, 1–15. Feldmann, R. J., and Maibach, H. I. (1967). Regional variation in percutaneous penetration of [14C] cortisol in man. J. Invest. Dermatol. 48, 181–183. Feldmann, R. J., and Maibach, H. I. (1974). Percutaneous penetration of some pesticides and herbicides in man. Toxicol. Appl. Pharmacol. 28(1), 126–132. Fitzgerald, R. H., and Manley-Harris, M. (2005). Laundering protocols for chlorpyrifos residue removal from pest control operators’ overalls. Bull. Environ. Contam. Toxicol. 75(1), 94–101. Guy, R. H., and Maibach, H. I. (1985). Calculations of body exposures from percutaneous absorption data. In “Percutaneous Absorption” (R. Bronaugh and H. I. Maibach, eds.), pp. 461–466. Dekker, New York. Guy, R. H., and Potts, R. O. (1992). Structure-permeability relationship in percutaneous pentration. J. Pharm. Sci. 81, 603–604. Hatch, K. K., and Maibach, H. I. (1986). Textile chemical finish dermatitis. Contact Dermatitis 12(1), 1–13. Hostynek, J., and Maibach, H. I. (2005). Advanced methods measure skin penetrants at the parts-per-billion level. Cosmet. Toiletries 120(11), 30–33. Istamboulie, G., Fournier, D. et al. (2009). Phosphotriesterase: a complementary tool for the selective detection of two organophosphate insecticides: chlorpyrifos and chlorfenvinfos. Talanta 77(5), 1627–1631. Knaak, J. B., Yee, K. et al. (1984). Percutaneous absorption and dermal dose cholinesterase response studies with parathion and carbaryl in the rat. Toxicol. Appl. Pharmacol. 76, 252–263. Loke, W.-E. et al. (1999). Wet decontamination-induced stratum corneum hydration – Effects on the skin barrier function to diethylmalonate. J. Appl. Toxicol. 19, 285–290. Lund, C. C., and Browder, N. C. (1944). The estimate of areas of burns. Surg. Gynecol. Obstet. 79, 61–70. Maibach, H. I. (1974). “Systemic Absorption of Pesticide through the Skin of Man”. Occupational Exposure to Pesticides: Federal Working Group. Maibach, H. I., Feldman, R. J. et al. (1971). Regional variation in percutaneous penetration in man. Pesticides. Arch. Environ. Health 23(3), 208–211. Marty, J. P. (1976). “Fixation des substances chemique dans les structures superficielles de la pesu: Importance les problemes de decontamination et de biodosponibilite,” Ph.D. Thesis. University of Paris-Sud, Paris. Moody, R. P., and Franklin, C. A. (1987). Percutaneous absorption of the insecticides fenitrothion and aminocarb in rats and monkeys. J. Toxicol. Environ. Health 20(1-2), 209–218. Moody, R. P., and Maibach, H. I. (2006). Skin decontamination: Importance of the wash-in effect. Food Chem. Toxicol. 44(11), 1783–1788.

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Moody, R. P., and Benoit, F. M. (1998). Dermal absorption of the insect repellent DEET in rats and monkeys: Effect of anatomic and multiple exposure. Personal communication. Nicoli, S., Guy, R. H. et al. (2008). Dermatopharmacokinetics: Factors influencing drug clearance from the stratum corneum. Pharm. Res. 26, 865–871. Obendorf, S. K., Csiszar, E. et al. (2003). Kinetic transport of pesticide from contaminated fabric through a model skin. Arch. Environ. Contam. Toxicol. 45(2), 283–288. Perkins, H. M., Rigakis, K. B. et al. (1996). The acceptability of chlorine bleach pre-treatment in the removal of chlorpyrifos residues from cotton and polyester/cotton fabrics. Arch. Environ. Contam. Toxicol. 30, 127–131. Qiao, G. L., Chang, S. K., and Riviere, J. E. (1993). Effects of anatomical site and occlusion of the percutaneous absorption and residue pattern of 2,6-[ring-14C] parathion in vivo in pigs. Toxicol. Appl. Pharmacol. 22(1), 131–138. Rougier, A., Dupuis, D. et al. (1986). Regional variation in percutaneous absorption in man: measurement by the stripping method. Arch. Dermatol. Res. 278(6), 465–469. Shah, P. V., Montoe, R. J., and Guthrie, F. E. (1983). Comparative penetration of insecticides in target and non-target species. Drug Chem. Toxicol. 6, 155–170. Stone, J. F., and Stahr, H. M. (1989). Pesticide residues in clothing. Case study of a midwestern farmer’s coverall contamination. J. Environ. Health 51, 273–276. Tickner, J., Friar, J. et al. (2005). The development of the EASE model. Ann. Occup. Hyg. 49(2), 103–110. Tregear, R. T. (1966). Molecular movement, the permeability of the skin. In “The Physical Functions of the Skin,” pp. 1–52. Academic Press, New York. U.S. Environmental Protection Agency (2003). “U.S. EPA bans ethyl parathion.” Available at  http://www.safer-world.org/e/chem/para/htm . Accessed October 11, 2009. Vallet, V., Cruz, C. et al. (2007). In vitro percutaneous penetration of organophosphorus compounds using full-thickness and split-thickness pig and human skin. Toxicol. in Vitro 21(6), 1182–1190. Van der Merwe, D., and Riviere, J. E. (2006). Cluster analysis of the dermal permeability and stratum corneum/solvent partitioning of ten chemicals in twenty-four chemical mixtures in porcine skin. Skin Pharmacol. Physiol. 19(4), 198–206. van Hemmen, J. J., Auffarth, J. et al. (2003). RISKOFDERM: Risk assessment of occupational dermal exposure to chemicals. An introduction to a series of papers on the development of a toolkit. Ann. Occup. Hyg. 47(8), 595–598. Van Rooy, T. G. M. et al. Absorption of polycyclic aromatic hydrocarbons through human skin: Differences between anatomic sites and individuals. J. Toxicol. Environ. Health, 38, 355–368. van Wendel de Joode, B., Vermeulen, R. et al. (2005). Accuracy of a semiquantitative method for Dermal Exposure Assessment (DREAM). Occup. Environ. Med. 62(9), 623–632. Wester, R. C., and Maibach, H. I. (1985). In vivo percutaneous absorption and decontamination of pesticides in humans. J. Toxicol. Environ. Health 16(1), 25–37. Wester, R. C., and Maibach, H. I. (1997). Toxicokinetics: Dermal exposure and absorption of toxicants. In “Comprehensive Toxicology” (J. Bond, ed.), pp. 99–114. Elsevier, Oxford. Wester, R. C., and Maibach, H. I. (1999). In vivo methods for percutaneous absorption and decontamination. In “Percutaneous Absorption” (R. Bronaugh and H. I. Maibach, eds.), pp. 215–227. Dekker, New York.

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Wester, R. C., Melendres, J. et al. (1991). Glyphosate skin binding, absorption, residual tissue distribution, and skin decontamination. Fundam. Appl. Toxicol. 16, 725–732. Westr, R. C., Melendres, J., and Maibach, H. I. (1992). In vivo percutaneous absorption and skin decontamination of alachlor in rhesus monkey. J. Toxicol. Environ. 36, 1–12. Wester, R. C., Noonan, P. K. et al. (1980). Variations in percutaneous absorption of testosterone in the rhesus monkey due to anatomic site of application and frequency of application. Arch. Dermatol. Res. 267(3), 229–235. Wester, R. C., Maibach, H. I. et al. (1984). In vivo percutaneous absorption of paraquat from hand, leg, and forearm of humans. J. Toxicol. Environ. Health 14(5-6), 759–762. Wester, R. C., Quan, D. et al. (1996). In vitro percutaneous absorption of model compounds glyphosate and malathion from cotton fabric into and through human skin. Food Chem. Toxicol. 34(8), 731–735.

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Wester, R. C., Hui, X. et al. (1999). In vivo skin decontamination of methy­lene bisphenyl isocyanate (MDI): soap and water ineffective compared to polypropylene glycol, polyglycol-based cleanser, and corn oil. Toxicol. Sci. 48(1), 1–4. Wiles, R., Hettenbach, T. et al. (1999). “Comments on the Preliminary Risk Assessment Document for Methyl Parathion,” Available at   http://www2.ewg.org/files/ewg_mp.pdf. Accessed October 11, 2009. Wu, C., and Chiu, H.-H. (2007). Rapid method for determining dermal exposures to pesticides by use of tape stripping and FTIR spectroscopy: A pilot study. J. Occup. Environ. Health. 36, 1–12. Zhai, H., Chan, H. P. et al. (2008). Skin decontamination of glyphosate from human skin in vitro. Food Chem. Toxicol. 46(6), 2258–2260.

Chapter 28

The Regulatory Evaluation of the Skin Effects of Pesticides Michael O’Malley University of California, Davis, California

28.1  Introduction 28.1.1  Basic Patterns of Skin Reaction Clinical effects of pesticides on the skin include both systemic and topical reactions. Systemic effects, such as urticaria, chloracne, and porphyria cutanea tarda, may occur following ingestion, inhalation, or topical exposure. Direct topical effects include acute irritation and corrosion, subacute (gradual-onset) irritation, and delayed-onset allergies. Any of the preceding injuries may damage the pigment-producing basal layer of the skin, resulting in either an increase or a decrease in epidermal melanin production. Typically, both injuries and residual effects occur in a pattern that coincides with the site of contact. Depending on the time interval between exposure and the onset of lesions, recognizing the source of the skin injury may be simple or complex. Distinguishing between allergic and irritant effects is a primary goal of both clinical and regulatory evaluation of the skin effect of pesticides. Clinically, irritant reactions tend to develop soon after exposure, whereas skin allergies are typically delayed in onset. Exceptions to this simple rule occur: Some irritant reactions are cumulative and some allergic reactions occur within minutes of contact with the offending allergen (urticaria). The following is a summary of the clinical protocol for provocations tests: Protocol for clinical patch testing Application of previously identified nonirritating concentration of test substance for 48 h, followed by removal of patch and initial reading. Follow-up reading at 96 h. Simplified scoring system for grading patch tests: 0 – no visible reaction 1 – erythema 2 – erythema and blistering 3 – necrotic reaction Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

In the regulatory arena, evaluation of the capacity of individual compounds to cause irritant or allergic reactions depends on animal testing as well as analysis of human use experience. Protocols for evaluating allergy and irritation in experimental animals are discussed next.

28.1.2  Testing Requirements and Test Protocols The requirements for skin testing of pesticides for the purposes of federal and state registration vary with the government jurisdiction. In the United States, primary dermal irritation studies and sensitization studies are required for each manufacturing-use product and each end-use product (U.S. Environmental Protection Agency (EPA), 1984). The tests performed in this manner are considered part of the regulatory database and are not available in the public literature. The irritation testing requirements are similar to those of the standard Draize tests. Dermal sensitization tests may be done according to one of several standard protocols.

28.1.2.1  Dermal Irritation Tests Irritation Test Using Albino Rabbits (Draize Test)  A single dose of the technical material with detailed characterization of contaminants, or an end-use product, is applied to the skin of one or several experimental animals (depending on the possibility of a corrosive reaction) for 4 h. Solid materials are moistened with distilled water, saline, or other vehicle at the time of application. The irritation is scored at intervals until the irritation has resolved or is considered permanent. Based on scores at 72 h and persistence of irritation for more than 14 days, materials are categorized as follows: Corrosive (category I; 72-h dermal irritation score 7) Severe irritants (category II; 72-h dermal irritation score 5–7) 701

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Moderate irritants (category III; 72-h dermal irritation score 2–5) Minimally irritant (category IV; 72-h dermal irritation score 0–2) – may be divided into compounds that produce no irritation (nonirritants) and those that produce transient, mild irritation Test limitations  As applied to most pesticide formulations, the Draize test does not distinguish between the effects of the active ingredient (AI) and the effects of carrier solvents or other components considered to be inert (as pesticides). As demonstrated with studies using 0.1% chlorothalonil (Flannigan and Tucker, 1985b), marked variations in irritation scores (vs. vehicle) occur depending on administration of the AI in saline (vehicle score  0, AI in vehicle score  0.04), petrolatum (vehicle score  0.33, AI in vehicle score  0.71), or acetone (vehicle score  0, AI in vehicle score  2.71). Another major limitation is the ordinal (rank) scoring system, which may be difficult to standardize between individual observers in the same laboratory and even more problematic to standardize between laboratories. The difficulties with test interpretation are illustrated by data in Section 28.2.5.11 on the herbicide oxadiazon: Two formulations contain 50% AI – a soluble powder that caused minimal irritation and a wettable powder reported to cause very severe irritation that did not reverse by the end of the study and was scored as corrosive. The information above on oxadiazon illustrates that testing of solid materials may be problematic in the Draize test. As discussed later, solids are inherently less likely than liquids to produce skin irritation. It is unclear to what extent moistening the test article prior to application compensates for this effect.

28.1.2.2  Sensitization Test Protocols In sensitization studies, following initial exposure to a test substance, the animals are challenged to establish whether they have developed hypersensitivity. This is evaluated by comparing scores during the induction period with those during the challenge period and with those of control animals that received the challenge without initial exposure. For ambiguous results on challenge, a rechallenge phase is used. Buehler Test (Closed Patch Method)  A closed patch is applied for 6 h, weekly, during a 3-week induction period; the test concentration for induction is chosen to be approximately 10-fold higher than the expected human exposure concentration. Challenge, with a nonirritating concentration, takes place during weeks 5–7. Open Epicutaneous Test  After establishing the concentration that produces minimal irritation and no irritation threshold, induction is begun at the latter concentration. Applications are repeated daily for 3 weeks or five times

Hayes’ Handbook of Pesticide Toxicology

weekly for 4 weeks, always on the same skin site. The challenge is conducted on day 21 using the minimal irritant and some lower concentrations; skin reactions are read after 24, 48, and/or 72 h. Rechallenge, if necessary, is done on day 35. Intradermal Methods Guinea Pig Maximization Test  Induction is begun on day 0 with 0.1 ml test material intradermal (by injection) together with 0.1 cc Freund’s complete adjuvant (FCA). Control animals receive only the injection of 0.1 ml PCA. On day 7, induction is boosted by occluding the test material against the skin for 48 h. On day 21, the challenge is performed on a shaved 4-cm2 area on the left flank using a nonirritating concentration of the test material. Controls are treated with occluded vehicle only. If challenge reactions are ambiguous, animals are rechallenged on day 28. Because the maximization test does not depend on dermal absorption during the induction phase (i.e., the test article is injected intradermally rather than being applied topically), the maximization test may demonstrate sensitization for compounds that are negative in the Buehler test (e.g., propargite, the type 2 pyrethroids cypermethrin and cyhalothrin, and the biological insecticides azadirachtin and abamectin). Freund’s Complete Adjuvant test  On induction days 1, 5, and 9, 0.1 ml of test material in PCA is injected into the shoulder of animals of the control group treated with FCA only. Challenge and rechallenge are performed on days 21 and 36, respectively: A minimum irritating concentration and a maximum nonirritating concentration are both tested in test animals and controls. On days 22–24, 36–38 skin sites are read 24, 28, and 72 h after challenge and rechallenge. The test is simple to perform and involves low material and operational expenses. Local Lymph Node Assay – A newer U.S. Environmental Protection Agency (EPA)-approved technique is the regional lymph node assay (LLNA) (Ashby et al., 1995; Ikarashi et al., 1994, 1996). The test involves use of a 3-day induction period, followed by monitoring of the uptake of H3labeled thymidine in regional lymph nodes (excised and placed in cell culture after sensitization) as a marker of sensitization. In addition to H3-labeled thymidine, sensitization can be monitored using cell number or levels of interleukin2 produced in cell culture (Hatao et al., 1995). In LLNA, the sensitization index (SI) provides a quantitative means of evaluating sensitization for a particular test solution by dividing the uptake of tritiated thymidine in treated animals by the uptake in vehicle-treated controls. A chemical is classified as a sensitizer if the following two criteria are fulfilled: (1) At least one concentration of the test chemical induces a stimulation index three times or greater than that of the vehicle control, and (2) the result must not be incompatible with a biological dose response.

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

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Table 28.1  Structural Basis of Skin Reactivity Related to Sensitization

Table 28.2  Physical and Chemical Properties Predictive of Skin Irritation

Protein reactivity

Physical or chemical property

IX IUNIQ ICOOR ICONJ IarOH IArNH IOH Transport/binding MR PL HBA HBD

Reactive aliphatic/aromatic halides Strong nucleophiles/reactive electrophiles Simple aliphatic esters Conjugated olefins: activated by SH/NH2 Michael addition Easily metabolized phenols of quinone Easily metabolized aniline precursor Oxidizable primary alcohols Polarizable molecular volume Lipophilic contribution of log P Hydrogen bond acceptor groups (count of electron pairs on N and O) Hydrogen bond donor groups (count of N–H and O–H)

Based on the SICRET model (Walker et al., 2005).

Chemicals with the lowest minimum sensitizing concentrations (EC3) are considered the most potent sensitizers in the assay. The lowest sensitizing concentration can be estimated from a limited number of test concentrations by employment of appropriate regression models (van Och et al., 2000). The results compare favorably to the more cumbersome 3to 7-week-long in vivo assays (Ikarashi et al., 1994), but a limited number of pesticide manufacturers have submitted LLNA data for purposes of pesticide registration. The vast majority of studies submitted to satisfy regulatory dermal sensitization data requirements have utilized the epicutaneous (Buehler) method. For interested readers, predictive skin testing procedures are described in more detail in Bashir and Maibach (2000).

28.1.2.3  Quantitative Structure–Activity Relationships The quantitative structure–activity relationship (QSAR) knowledge-based computer system DEREK was developed for predicting skin allergens under the collaboration of a multinational group of toxicologists (Enslein et al., 1997; Hostýnek, 1998; Hostýnek et al., 1996; Magee et al., 1994a,b,c; Ridings et al., 1996; Sanderson and Earnshaw, 1991). The individual chemical moieties or structural elements related to sensitization and irritation are presented in Tables 28.1 and 28.2. Limitations of the DEREK model have been discussed by Fedorowicz et al. (2005), who found that electrophilic interactions explain less than half of allergens recognized to cause delayed contact sensitivity. In this chapter, the model is used to help clarify common structural ­properties and understand skin effects in terms of their chemical ­ reactivity and

No. passed/ Effect tested

All chemicals Melting point 200°C Log Pow or log Kow 3.1 Lipid solubility 0.01 g/kg

291/297a 56/56 60/60

No I or C No I or C No C

Group C (CxHyOz) Melting point 55°C Molecular weight 350 g/Mol Surface tension 62 mN/m Vapor pressure 0.0001 Pac4

128/130a 93/93 94/95b 73/73

No I or C No C No C No I

Group CN (CxHyOzNa) Lipid solubility 0.4 g/kg Molecular weight 290 g/Mol Aqueous solubility 0.1 g/l Log Pow or log Kow   4.5 Vapor pressure 0.001 Pa Molecular weight 540 g/Mol Melting point 180ºC Aqueous solubility 0.0001 g/l Log Pow or log Kow   5.5

56/56 338/338 280/280 119/119 273/273 86/86 153/153 104/104 85/85

No I or C No C No C No C No C No I No I No I No I

Group CNHal (CxHyOzNaF, Cl, Br, or I) log Pow or log Kow   3.8 Aqueous solubility 0.1 g/l Molecular weight 370 g/Mol Lipid solubility 400 g/kg Molecular weight 380 g/Mol Lipid solubility 4 g/kg Aqueous solubility 0.001 g/l

70/70 135/135 109/109 76/76 99/99 29/29 78/78

No I or C No C No C No C No I No I No I

Group CNS (CxHyOzNaSb) Molecular weight 620 g/Mol Melting point 50ºC Surface tension 62 mN/m Melting point 120ºC Log Pow or log Kow   0.5

53/53 179/180a 92/92 137/137 96/96

No C No C No C No I No I

Group CHal (CxHyOzF, Cl, Br, or I) Molecular weight 370 g/Mol Molecular weight 280 g/Mol

24/24 59/59

No I or C No C

C, skin corrosion; I, skin irritation. Based on the SICRET model (Walker et al., 2005). a Chemicals that did not pass were organic salts, which release strong inorganic acids or bases when in contact with aqueous substrates/ organic media. b Chemical that did not pass was a skin defatting ether with high vapor pressure at 20°C. c The model uses the SI unit Pascal (Pa) for pressure, whereas the data in the text use mm Hg: 133 Pa  1 mm Hg. For large biological compounds, the molecular weight is expressed in Daltons (Da), a unit equivalent to 1 g/Mol; kilodaltons (kDa); or megadaltons (MDa).

their biologic mode of action. The isocyanate compounds, for example, are reactive nucleophiles that act as nonspecific enzyme inhibitors (Roberts et al., 1998). Binding at the heme iron (Fe2) in fungal cytochromes, for example, would be expected to inhibit cellular respiration. Binding at other

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704

sites may produce allergic reactions. Data on methyl isothiocyanate (MITC)-generating fumigants and thiocyanate-containing fungicides are discussed later. A model for the prediction of skin irritation and corrosion [Skin Irritation Corrosion Rules Estimation Tool (SICRET)] has also been described (Walker et al., 2005) that contains elements that overlap with the predictive skin sensitization model. A list of negative predictors based on physical properties is shown in Table 28.2, and positive identifiers based on the presence of chemical groups or moieties are shown in Table 28.3. Because of the importance of the physical properties in predicting their effects on the skin, physical–chemical data for individual compounds are included in this chapter. Except as noted, data were obtained from the National Library of Medicine website TOXNET (National Library of Medicine, 2009). Predictive models such as DEREK and SICRET have not yet replaced sensitization and irritation testing in animal models. Nevertheless, the structural moieties identified in the models provide an important tool for understanding the otherwise confusing list of pesticides that cause irritation and sensitization. In some cases, the reactive chemical elements correlate with other biological types of activity. For example, many fungicides are strong electrophiles. As suggested by the “reactive electrophile” hypothesis (Guengerich, 2001; Miller and Miller, 1981), many are also animal carcinogens (O’Malley, 2007). As discussed later, many are reported skin irritants or skin sensitizers.

Table 28.3  Chemical Moieties Predictive of Skin Irritation Chemical group or moiety/structural class

The principal regulatory decision dependent on the results of preregistration animal testing concerns the content of precautionary statements on the pesticide product label. Materials showing corrosive effects or reversible skin irritation are labeled as such. Labels for products not found to be corrosive or irritating generally carry statements advising the user to minimize the degree of skin contact. Materials judged as sensitizers in animal tests or reported as sensitizers in the public domain scientific literature are required to indicate the possibility of skin sensitization on the product label.

28.1.4  Integration of Illness Surveillance Data with Experimental Dermal Irritation and Sensitization Tests Apart from tests required for registration, additional regulatory information is obtained from postregistration surveillance of pesticide illness reports and from public literature reports on adverse skin effects of pesticides. A summary of information on irritation, sensitization, and postregistration illness surveillance information derived from

All chemicals -Alkynes

I

Group C (CxHyOz) Acrylic acids Ortho- and para-quinones

C C

Acids, including aliphatic saturated acids, halogenated acids Aldehydes Phenols Catechols, resorcinols, hydroquinones Catechols, resorcinols, hydroquinones – precursors Acid anhydrides Ketenes -Lactones Lactones Epoxides Acrylic and methacrylic esters Ketones C10–C20 Aliphatic alcohols Ethyleneglycolethers (Hydro)peroxides

28.1.3  Regulatory Decisions

Potential effect

I or C I or C I or C I or C I or C I or C I or C I or C I or C I or C I I I I I

Group CN (CxHyOzNa) Quaternary organic ammonium/phosphonium salts Di/trinitrobenzenes Alkylalkanolamines -Lactams Acid imides Aromatic amines

C C I or C I or C I or C I

Group CNHal (CxHyOzNaF, Cl, Br, or I) Carbamoyl halides Halonitrobenzenes

I or C I or C

Group CNS (CxHyOzNaSb) -Halogenated amides, thioamides

I

Group CHal (CxHyOzF, Cl, Br, or I ) Benzyl halides Halogenated alkanes and alkenes Tri- and tetrahalogenated benzenes

I or C I I

C, skin corrosion; I, skin irritation. Based on the SICRET model (Walker et al., 2005).

California Department of Pesticide Regulation (CDPR) data for individual AIs, by use and structural category, is shown in Table 28.4. The data are based on review of registration data on dermal irritation and sensitization by the CDPR medical toxicology program since the mid 1990s.1 Probable or definite skin illness or injury cases involving the same products are also shown in Table 28.6. Because the pesticide illness database (1982–2006) contained more than 1

The previous version of the chapter was based on summary memoranda produced between January 1989 and July 1997.

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

705

Table 28.4  Summary of Irritation, Sensitization, and Postregistration Illness-Surveillance Data Derived from the California Department of Pesticide Regulation Use category

Chemical structure category (No. of cases reported 1982–2006)

Mixer/ loader

Applicator

Total cases

% of all handler cases

1

1

0.1

5

19

24

1.2

Acaricide Adjuvant Disinfectant

Organic peroxides (7) Pine tar distillate (12) Isothiazolines (21) Aldehydes (53) Triazine chlorine releasers (85) Phenols (56) Quaternary ammonium (274) Inorganic halogen (472)

101

727

828

42.4

Fumigant

Thiadiazine (2) Ethylene epoxide (3) Inorganic (5) Phosphide (6) Thiocarbamate (59) Halogenated hydrocarbon (99)

20

154

174

8.7

Fungicide

Chlorophenol (5) Imidazolidine (5) Triazole (9) Carbamate (14) Phthalimido (15) Thiocarbamate (16) Benzonitrile (18) Triazine (27) Coal tar mixture (37) Inorganic (105)

33

232

265

13.5

Herbicide

Triazole (5) Thiocarbamate (5) Cyclohexene (7) Dinitrobenzenesulfonamide(7) Petroleum distillate (7) Phenoxy (14) Phenoxy-trifluoro-pyridine (14) Nitrophenol (17) Dinitro-trifluoromethyl (18) Bipyridyl (57) Amino acid phosphonate (204)

37

347

384

19.3

3

5

8

0.4

76

216

292

14.7

1 8 2 1 1713

1 10 2 1 1990

0.1 0.5 0.1 0.1 100

Insect growth regulator Insecticide

Piscicide Plant growth regulator Repellant Rodenticide Total

Inorganic (8) Biological (9) Organometallic (11) Organochlorine (12) Carbamate (17) Pyrethroid (40) Organophosphate (77) Propargyl derivative (105)

2

277

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706

19,000 reports of skin injury or illness from both agricultural and nonagricultural use, the review was limited to 1990 cases involving possible, probable, or definite skin effects to pesticide applicators from a single AI (see Table 28.4). Few of the 1990 cases had patch testing performed. For purposes of this review, these cases are referred to in the text as the “pesticide handler database” or “handler database.”

28.2  Review of use categories 28.2.1  Antimicrobial Agents/Disfinfectants Many of the antimicrobial agents registered as pesticides are corrosive or markedly irritant. This suggests that antimicrobial activity may be correlated with underlying chemical reactivity necessary to provoke sensitization. Mode of Action  Isothiazolin compounds, aldehydes, copper compounds, carbamates, ozone, and peroxides act as electrophilic enzyme inhibitors. Membrane-active microb­ ials include quaternary ammoniums, biguanides, phenols, and alcohols (Williams, 2007). California data on pounds sold for 2007 indicate the wide usage of these compounds (Table 28.5). Illness Data  The applicator database contained 828 total cases involving antimicrobial agents/disinfectants. By chemical category, this included cases caused by organic peroxides, pine tar distillate, isothiazolines, aldehydes, triazine chlorine releasers, phenols, and quaternary ammonium compounds (see Table 28.4).

Table 28.5  Amount of Antimicrobial Agents/ Disinfectants Sold in California, 2007 Compound

Pounds sold in California, 2007

1,2-Benzisothiazolin-3-one

2,213,096

2-Methyl-4-isothiazolin-3-one

564,767

5-Chloro-2-methyl-4-isothiazolin-3-one

1,332,803

Octhilinone

2,543,337

Ca hypochlorite

7,716,557

Na hypochlorite

192,432,794

Chlorine

83,500,886

Dichloro-s-triazinetrione

1,945

Sodium dichloro-s-triazinetrione

5,871,688

Sodium dichloro-s-triazinetrione dihydrate 6,371,099 Quaternary ammonium compounds

3,797,937

28.2.1.1  Isothiazolins (Kathon Compounds) O

O

N

N

S

S 2-Methyl-4-isothiazolin-3-one O

Cl 5-Chloro-2-methyl-4-isothiazoline O

N

N

S

S

1,2-Benzisothiazol-3(2H)-one

Octhilinone

Isothiazolin compounds

Physical Properties  1,2-Benzisothiazolin-3-one: molecular weight (MW), 151.1826; vapor pressure (VP), NA; melting point (MP), 100°C, 152°C; solubility, miscible in water; oil VP, NA; log P, NA 2-Methyl-4-isothiazolin-3-one: MW, 115.1496; MP, 50–51°C; solubility, 30 g/l H2O, miscible in organic solvents; VP, 0.025 kPa (0.1875 mm Hg); log P, NA 5-Chloro-2-methyl-4-isothiazilone-3-one: MW, 149.5947; MP, 52–55°C; VP, NA; log P, NA Octhilinone: MW, 213.3372; boiling point (BP), 120°C; water solubility, 0.5 g/l; VP, NA; log P, NA Irritation Data  46.5, 27.5, 5.0% concentrations of octhilinone (2-n-octyl-4-isothiazolin-3-one) are all corrosive in the Draize irritation assay. Assays reported for 1,2-benzisothiazolin-3-one show some inconsistency: 8.5 and 11.14% formulations were reported to cause only transient irritation (Draize category IV), whereas 9.5 and 20% formulations were reported to be corrosive (Draize category I). A 5% formulation of 2-methyl-4,5-trimethylene-4-isothiazolin-3one was on two occasions evaluated with the Draize assay: Severe edema was seen during the first 2 or 3 days after dosing. This edema regressed, disappearing by day 14 in all animals (Draize category II). A second test showed mild erythema and edema that disappeared by 72 h, with a treatment-related eschar that persisted for 3 or 4 days in one of the six animals tested (Draize category III). Sensitization Data  In sensitization assays, a formulation with 20% 1,2-benzisothiazolin-3-one was positive (SI 3 for 20%, 50% dilution of test article) in the local lymph node assay; a 9.5% formulation and a 19.8% formulation were both negative in the Buehler assay. The 19.8% formulation was also negative in the guinea pig maximization test (GPMT). There were no sensitization studies reported for octhilinone. Numerous reports in the public literature document their capacity to sensitize (Bruze and Gruvberger, 1988; Emmett et al., 1989; Foussereau et al., 1984; Mathias et al., 1983; Menne, 1991; Menne et al., 1991; Pilger et al., 1986; Thormann, 1982). In the pesticide handler database, the

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

i­sothiazolin ­ compounds appeared to be relatively frequent causes of irritant skin reactions or frank chemical burns following accidental direct contact (21 total cases; individual examples include cases 89-1312 and 90-552, listed in Table 28.6). The data were limited, however, because the individual isothiazolin compounds involved in each case were not identified.

28.2.1.2  Quaternary Amines Physical Properties  Properties not available for these complex mixtures. Irritation Data  Quaternary ammonium derivatives cause irritation in the Draize assay in higher concentrations (e.g., more than 5%), as demonstrated for the four compounds listed. Sensitization Data  Data on sensitization for quaternary ammonium compounds are less clear-cut than the irritation data. Predictive testing for registered formulations containing mixtures of typical compounds with C10–C20 alkyl substituents were usually negative [e.g., a formulation with dioctyl dimethyl ammonium chloride, didecyl dimethyl ammonium chloride, octyl decyl dimethyl ammonium chloride, and alkyl (50% C14, 40% C12, and 10% C16) dimethylbenzyl ammonium chloride, tested by the Buehler method]. However, individual quaternary ammonium compounds with epoxy moieties or other reactive substituents proved to be sensitizers, as reported in the public domain literature (Estlander et al., 1997; Kanerva et al., 2001). Illness Data  Cases related to the quaternary ammonium compounds in the handler database involve mixtures and are grouped together in the handler database: The 274 cases associated with quaternary amines accounted for 13.8% of the total number reported, with typical cases (e.g., case 1988-1893, see Table 28.6) occurring on direct contact.

28.2.1.3  Chlorine Compounds and Triazine Chlorine Stabilizers Sodium hypochlorite Na

OH N

H N H5C2

N N

H

OH

O

N N

H hexahydro-1,3,5-triethyl-s-triazine

Irritation Data  Three hypochlorite products (2–6% AI) caused severe irritation in the Draize assay. Two products (1–2% AI) caused moderate irritation. Five formulations (0.0164–4% AI) caused mild irritation. Cyanuric acid was corrosive or severely irritant in the Draize assay (study concentrations 25%) (see Table 28.6). The chlorine stabilizer hexahydro-1,3,5-triethyl-s-triazine was corrosive in the Draize test, as was the related chlorinated product (sodium dichloro-s-triazinetrione dihydrate). Sensitization Data  Sensitization to hypochlorite has been reported in the form of contact urticaria (Hostýnek et al., 1989). Nevertheless, most sodium hypochlorite-containing antimicrobials are not labeled as sensitizers. Although testing with the Buehler and guinea pig maximization protocols did not show evidence of sensitization (see Table 28.6), sodium dichloro-s-triazinetrione dehydrate caused asthma and urticaria among members of a British ophthalmology practice (Goverdhan and Gaston, 2003). No sensitization study on cyanuric acid was available for review, but a 95% liquid product containing the chlorine stabilizer hexahydro-1,3,5-triethyl-s-triazine was weakly positive in the Buehler assay. Illness Data  Chlorine, inorganic chlorine salts (hypochlorites), and triazine chlorine stabilizers accounted for 557 (28.0%) of the cases in the handler database. The cases reported in the handler database occurred principally in end users of sanitizers and disinfectants. These typically occurred on direct contact (e.g., see Table 28.6, case 1987-1468).

N

H2O

HO

Cl

O− Na+

CH3

O

N N

OH

Cyanuric acid C2H5

Physical Properties  Sodium hypochlorite physical properties: formula, ClHO.Na; MW, 74.44; MP, NA; VP, NA; solubility in H2O, 29.3 g/l Cyanuric acid physical properties: formula, C3H3N3O3; MW, 129.08; MP, 360°C; log P, 1.95; VP, NA; solubility in H2O, 2593 mg/l at 25°C; other solubilities, slightly soluble in common organic solvents such as acetone, benzene, diethyl ether, ethanol, and hexane Hexahydro-1,3,5-triethyl-s-triazine physical properties: formula, C9H21N3; MW, 171.28; BP, 207–208°C; VP, NA; log P, NA; solubility in H2O, NA Sodium dichloro-s-triazinetrione dihydrate physical properties: formula, C3HCl2N3O3.Na.2H2O; MW, 220.96; VP, NA; log P, NA; solubility in H2O, 227 g/l

Cl

OCl

C2H5

707

H2O

Sodium dichloroisocyanurate dihydrate

HO ortho-phenylphenol

C

CH3

CH3

para-tert-butyl phenol

708

Table 28.6  Animal Testing and Illness Data Compound or group Identifiers Compound

Predictive Tests in Animals CAS #

Draize irritation test

Data from pesticide handler data base 1982–2006 Sensitivity

Isothiazolin (Kathon compounds) 1,2-benzisothiazolin3-one

2634-33-5

2-methyl-4-isothiazolin- 2682-20-4 3-one

# of cases Case examples 21

4 formulations (9.5% liquid–19.5% liquid) caused corrosion. 4 formulations (5.1% liquid, 1 5.1% and 11.52% mixture, and a 19.8% liquid) caused moderate irritation. 3 formulations (8.5% liquid, 10.9% mixture with other antimicrobials, 72% paste) caused minimal irritation in the Draize assay.

Negative in Buehler, GPMT; positive LLNA; Sensitizer per public domain literature

97.8% melted solid and 50.4% liquid corrosive; mixture with total 5% isothiazolin minimal irritant

Sensitizer per public domain literature

2-Methyl-4,582633-79-2 4.85% liquid severe irritation; 95% solid Trimethylene-4moderate irritation Isothiazolin-3-one (MTI)

1989–1312: A worker sanitizing cooling towers with a Kathon® compound accidentally spilled some of the solution on himself and suffered a chemical burn. 90–552: An employee adding concentrate (of Kathon®) to corrosive domain literature washing solution spilled some of the material onto trouser leg of work pants, causing a chemical burn.

Sensitizer per public domain literature

2-n-octyl-4-isothiazolin- 26530-20-1 27.5%, 45% and 46.5% liquids corrosive; Sensitizer per public 3-one 4.85% liquid severe irritation domain literature Quaternary ammonium compounds

68424-85-1 50% liquid corrosive

alkyl dimethyl benzyl 112-18-9 ammonium chloride (multiple compounds alkyl groups  c14,c16 etc.)

11 liquid products (1.19%–80%AI) caused corrosion in the Draize assay.

6 products (0.3%–10% liquid) caused moderate irritation. 5 products (0.25%–2.0% liquid) caused minimal irritation.

No dermal sensitization study available

Nonsensitizer by Buehler tests (6 products 0.3%–80% AO)

1988–1893: A worker splashed a sanitizer containing quaternaryammonium chloride concentration modified Maguire ammonium compounds onto his face. Four hours after exposure he developed 6–10 macular lesions at the site where the material splashed on him.

Hayes’ Handbook of Pesticide Toxicology

alkyl (60%C14, 25%C12, 15%C16) dimethyl benzyl ammonium chloride

274

Table 28.6  (Continued) Compound or group Identifiers

Predictive Tests in Animals

Data from pesticide handler data base 1982–2006

CAS #

Draize irritation test

Sensitivity

dioctyl dimethyl ammonium chloride

5538-94-3

0.5% mixed quat formulation severe irritant

Nonsensitizer by Buehler test (0.5% mixed quat product)

trimethyl ammonium chloride (didecyl dimethyl ammonium chloride)

7173-51-5

11 formulations (3.85%–50% liquid) caused corrosion in the Draize assay. A 1% liquid caused severe irritation and a 4.23% liquid moderate irritation. 3 products(0.13%–4.23% liquid) caused minimal irritation.

Nonsensitizer

# of cases Case examples

Chlorine compounds and chlorine stabilizers – Cases reported represent irritant reactions in end-users of sanitizers/disinfectants

557

cyanuric acid

108-80-5

Two formulations (a 96% powder and No data a 99% powder) caused corrosion in the Draize assay. 3 products (each with 25% powder) caused moderate irritation and a 6% powder caused minimal irritation.

54

1991–2327: A custodial employee splashed material on her right arm while cleaning a toilet and subsequently developed itchy, red, and swollen area at the site of contact.

hexahydro-1,3,5triethyl-s-triazine

7779-27-3

96.3% liquid caused severe irritation, borderline corrosion

472

1987–1468: A pet store employee developed a severe, painful rash on her hands from using a 12.5% sodium hypochlorite product to clean kennels without wearing gloves. The product proved to be a swimming pool sanitizer used in violation of the product.

56

1988–909: A hospital janitor got disinfectant on the hand through a hole in disposable glove, caused burning of the skin, diagnosed as irritant contact dermatitis.

Sodium dichloro-s51580-86-0 triazinetrione dihydrate

sodium hypochlorite

7681-52-9

95% liquid weakly positive in Buehler assay 99.4% solid GPMT negative; 1.5%Copper sulfate pentahydrate, 93.5% trichloro-striazinetrione Buehler negative

Three hypochlorite products (2%–6% AI) caused severe irritation in the Draize assay. Two products (1–2% AI) caused moderate irritation. Five formulations (0.0164%–4% AI) caused mild irritation

Two formulations (0.6%–2.4% AI) were non-sensitizers in the Buehler assay, but reported as a sensitizer in the public domain literature.

Phenolic compounds: 128 phenolic compounds registered for use as disinfectants; typical examples include ortho-phenylphenol, p-tert-butyl phenol o-phenylphenol

90-43-7

11% formulation severe irritant; mixed formulation 5% of 5% of ortho-benzylpara-chlorophenol, 10.5% of orthophenylphenol, corrosive

Negative in GPMT

p-tert-butylphenol

98-54-4

No data

Sensitizer per public domain literature

709

(Continued )

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

Compound

710

Table 28.6  (Continued) Compound or group Identifiers Compound

Predictive Tests in Animals

Data from pesticide handler data base 1982–2006

CAS #

Draize irritation test

Sensitivity

# of cases Case examples

56-35-9

1.06% tributyltin oxide, 1.6% 5-chlor-2(2,4-di-chlorophenoxy) phenol severe irritant; 0.3% tributyltin oxide, 0.7% chlorothalonil moderate irritant

No data

1

Organotins tributyltin oxide

tributyltin methacrylate, 2155-70-6; 3 products (a mixture of 33.8% copper tributyltin benzoate and 4342-36-3; oxide and 8.3% tributyltin methacrylate, tributyltin fluoride 1983-10-4 4% tributyltin fluoride and 10.7% tributyltin methacrylate, and 30.6% tributyltin methacrylate) caused corrosion in the Draize assay.

21.2% alkylamine HCL, 0 15% tributyltin benzoate product and 8.3% tributyl tin methacrylate product both sensitizers in Buehler test

Miscellaneous antimicrobial compounds 2,2-dibromo-3-nitrilpropionamide

10222-01-2 20% liquid corrosive in Draize test; 98.2% solid severe irritant; 40% solid minimal irritant

46 0

2-(hydro34375-28-5 Labeled as a corrosive xymethylamino)-ethanol

product with 50% AI nonsensitizer in GPMT; formaldehyde breakdown product a dermal sensitizer

0

Bronopol

52-51-7

Sensitizer per public domain literature

0

1,2-dibromo-2, 4-dicyanobutane

35691-65-7 98% wp applied as 50% mixture with mazola oil, caused moderate-severe irritation not reversible at 72 hours

Sensitizer per public domain literature

0

Glutaraldehyde

111-30-8

3 products 10%–51.3% glutaraldehyde Sensitizer per public caused corrosion in the Draize assay. domain literature A mixture of 14% glutaraldehyde and 2.5% quaternary ammonium caused severe irritation. 4 products (0.24%–3.2% glutaraldehyde) caused minimal irritation

Iodine

7553-56-2

99.5% technical material caused corrosion; other products (0.6%–46.9% ai) caused minimal irritation

45

46.9% liquid product 1 sensitizer in Buehler test; 0.6% liquid product negative

87–2288: While disinfecting with an iodine product, an employee developed pruritic rash on arm.

Hayes’ Handbook of Pesticide Toxicology

20% liquid sensitizer in Buehler test

18.2% concentrate severe irritant in Draize assay; mixtures with isothiazolin corrosive

1990–728: An employee added a mildewcide to a can of paint. When she pounded the lid back on the paint can, some of the material splashed on her, resulting in a rash on the face and neck.

Table 28.6  (Continued) Compound or group Identifiers CAS #

Draize irritation test

Data from pesticide handler data base 1982–2006

Sensitivity

# of cases Case examples

Insecticides Organophosphates acephate

63 30560-19-1 5 products containing acephate powder 3 products (1.5%– 5 or pellets (1.5%–90% AI) caused minimal 97.4% AI) were nonirritation in the Draize assay. sensitizers in the Buehler assay.

89–2500: An applicator developed a rash, described as urticaria and contact dermatitis, on both arms soon after acephate. The symptoms disappeared soon after he showered, but reappeared when he next applied acephate. 83–2409: An applicator spraying acephate on trees was exposed to liquid insecticide soaking through his clothes from a leaking fitting which allowed material to soak through his clothes; developed contact dermatitis. 86–1084: A structural pest control operator was trying to attach a crack and crevice injector to a spray can containing acephate. He sprayed his face and hands, and contaminated the respirator he was wearing and subsequently developed erythematous papules and vesicles on forearms, hands, and ears.

chlorpyrifos

2921-88-2

Technical chlorpyrifos (97.6% AI) caused transient irritation; some EC formulations with 40% AI caused moderate-severe irritation; dilute formulations with   1% AI all caused minimal irritation.

Two EC formulations (22.8% and 24.66% AI) caused sensitization in the Buehler Assay and two (42.8% and 44.9% AI) did not. Both products tested (41% EC and 30% EC, also containing 0.54% gamma-cyhalothrin) were sensitizers in the LLNA.

23

coumaphos

56-72-4

Technical coumaphos (98.25% AI ), a formulation with 25% wettable powder, and liquid formulations of 12.1% and 11.9% all caused minimal irritation in the Draize assay.

Possible sensitizer in Buehler assay in 11.9% EC, other products (98.25% powder and 25% WP) negative

0

Diazinon

333-41-5

Four products (47.5%–48% EC, AC, or FC) caused moderate irritation in the Draize assay. The 87% technical material and 18 other formulation (0.5%–87%AI) caused minimal irritation. Three products were associated with severe irritation or corrosion (0.5%–0.58% AI)

Sensitizer in GPMT per 10 public domain literature; multiple formulations negative in Buehler assay

1985–343: An applicator treating a large carpet area for fleas came using a backpack, noticed that some of the spray material (chlorpyrifos) had leaked, soaking his lower back and upper Iegs. He had burning at the site of contact, but did not develop overt dermatitis.

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

Compound

Predictive Tests in Animals

1987–2537: An applicator developed a rash on his arms and chest after a hose ruptured. He changed his shirt, but did not shower.

(Continued )

711

712

Table 28.6  (Continued) Compound or group Identifiers

Predictive Tests in Animals

Data from pesticide handler data base 1982–2006

CAS #

Draize irritation test

dichlorvos (DDVP)

62-73-7

1% flybait causes minimal irritation in the No data Draize assay

1

1983–1870: An apartment manager was spraying dichlorvos for roach control when the hose broke on his hand sprayer, splashing material on his right arm and left ear. On exam he had severe dermatitis of the right elbow and forearm and left ear, complicated by a possible secondary infection.

dimethoate

60-51-5

Technical material caused minimal irritation; a 32.7% EC caused moderate irritation

Buehler sensitization studies negative (32.7% EC and 44.7% EC), but technically limited because of absence of positive controls

3

1983–1880: Even though protective clothing was worn, applicator developed a rash after application of dimethoate on grapes. He also had nonspecific symptoms of systemic poisoning (nausea and headache).

fosthiazate

98886-44-3 48.4% EC reported to cause severe 48.4% EC sensitizer in irritation in one lab, but minimal irritation GPMT, but negative in in another; 10% granular product and the Buehler assay 75% EC both caused minimal irritation

0

malathion

121-75-5

10

1994–401: An employee of a small central valley city was pumping up a spray tank containing malathion when a hose coupling broke, spraying the material on his face and neck. Despite washing immediately, he developed a mild erythema in the exposed areas.

methamidophos

10265-92-6 40% EC is a minimal irritant in the Draize 40% EC is a assay nonsensitizer in Buehler test

1

1984–129: A mixer/loader splashed mixture of methamidophos and buffer on himself while transferring material and developed blisters in the exposed area.

methidathion

950-37-8

22.6% EC causes moderate irritation in Draize assay

25% formulation nonsensitizer in Buehler test

3

No cases of dermatitis after direct exposure

Naled

300-76-5

87% liquid and 78% EC formulations Sensitizer according to corrosive in Draize assay; liquid technical public domain literature material (94.5% AI) moderately irritating

7

1988–942: A worker hand poured naled (Dibrom®) for application on strawberries, without wearing rubber boots, gloves, respirator, or eye protection. After spilling the material on his leather boots, he wore them the rest of the day. He developed severe blister on foot, which did not improve with home treatment and eventually required medical attention. 88–2330: A worker was using naled for fly control, when a hose broke, spraying him in the face. He developed a rash on the ears despite wearing coveralls, gloves, respirator, and goggles. At the time of treatment, he was noted to have a chemical contact dermatitis with a secondary infection.

Technical material causes transient, minimal irritation

Sensitivity

Nonsensitizer in animal studies: 20% product labeled as sensitize

# of cases Case examples

Hayes’ Handbook of Pesticide Toxicology

Compound

Table 28.6  (Continued) Compound or group Identifiers

Predictive Tests in Animals

Data from pesticide handler data base 1982–2006

CAS #

Draize irritation test

Sensitivity

oxydemeton-methyl

302-12-2

no data

A 50% formulation of 2 oxydemeton-methyl was a nonsensitizer in the Buehler assay.

1983–304: A worker was cleaning the tip of a spray rig when oxydemeton-methyl (Metasystox®) splashed into face underneath the shield he was wearing. He developed erythema at the site of contact.

parathion and methyl parathion

56-38-2

98% liquid technical material caused no irritation in the Draize assay

Possible cases of sensitization reported in public domain literature

1

1985–277: While loading parathion, some splashed down rig onto cloth coveralls. The affected worker developed a rash on the groin.

phosmet

732-11-6

70% wettable powder causes transient irritation; 11.6% EC causes moderate irritation

5% dust is a nonsensitizer in Buehler test

5

1986–318: A San Diego pet shop employee developed a rash on her hands after she began using a phosmet flea dip. 1989–1494: Washing her dog with insecticidal shampoo when she developed redness and numbness of the hands.

tetrachlorvinphos

22248-79-9 4 formulations (1.08% spray–99% 99% powder sensitizer powdered AI) caused minimal irritation in in Buehler test the Draize assay

Carbamates

# of cases Case examples

0

17

aldicarb

116-06-3

92% technical liquid causes severe systemic toxicity without causing irritation; a 15.87% granular product caused no irritation or systemic toxicity

1

No cases of dermatitis following direct accidental contact; high systemic toxicity on skin contact

bendiocarb

22781-23-3 Reported as minimal irritant in public domain literature

Mixture with 2.5% 2 bendiocarb and 12% PBO is a nonsensitizer in the Buehler test

1982–1278: While treating for cockroaches with bendiocarb, a hotel employee got his fingers into the material. He later stuck his fingers in his mouth, causing a condition described as a mild allergic reaction to the lips and tongue.

carbofuran

1563-66-2

Reported as minimal irritant in public domain literature

1

No cases of dermatitis following direct accidental contact; high systemic toxicity on skin contact

carbaryl

63-25-2

99% technical material, granular products Nonsensitizer in Buehler 6 (6.3%–7% AI) caused minimal irritation in test 90% DF product the Draize assay; 0.5% shampoos caused moderate irritation

1982–2634: A turkey farm employee developed dermatitis after a hose broke during an application of carbaryl. 1982–2703: An applicator applying carbaryl dust developed dermatitis after getting the material on his hands and arms.

fenoxycarb

72490-01-8 40% wettable powder caused severe Nonsensitizer in Buehler 0 irritation in Draize assay; technical test material, 23% liquid, 1.2% aerosol, 1% granular ant bait caused minimal irritation

methiocarb

2032-65-7

75% concentrate, 2% pellet, and 1% aerosol caused minimal irritation in the Draize assay.

No data available

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

Compound

Nonsensitizer in Buehler 0 test

(Continued )

713

714

Table 28.6  (Continued) Compound or group Identifiers

Predictive Tests in Animals

Compound

CAS #

methomyl

Draize irritation test

Data from pesticide handler data base 1982–2006

Sensitivity

# of cases Case examples

16752-77-5 98.4% technical solid and 92.4% powdered concentrate caused minimal irritation in the Draize assay nonirritant

1% methomyl/0.025% muscalure negative in the Buehler test; 98.4% technical solid negative in local lymph node assay

2

1983–285: Failing to close cap, the sprayer leaked on his pant leg, irritating an open sore. 1984–1512: While spraying methomyl on corn, an applicator cleaned clogged nozzles on his equipment with his bare hands and developed a bad rash.

oxamyl

23135-22-0 42% liquid formulation caused minimal irritation in the Draize assay

10% liquid positive in LLNA

1

No cases of dermatitis following direct accidental contact

propoxur

114-26-1

Technical material nonirritant

1% RTU product nonsensitizer in Buehler test

4

1988–297: A structural pest control worker was spraying propoxur, and wiped his hands on shirt. He then developed rash on chest where he wiped his hands.

pyrethrins

121-21-1

57% technical material causes transient irritation

Five products (0.124%– 9% pyrethrins, mixed with other insecticides and synergists) caused sensitization in the Buehler test.

40

Pyrethrin cases all involved mixtures that did not meet the case definition for the the pesticide handler database, examples include: 1990–2621: A fairgrounds employee suffered chemical burn to right leg while applying a pyrethrin insecticide to livestock barns. “The fogger” machine he was using had a loose cap on the reservoir tank causing insecticide concentrate to come in contact with his leg. 1986–385: A kitchen employee set off a fogger and remained in the treated area for 10 minutes in violation of the label. He developed skin irritation on his face.

piperonyl butoxide

51-03-6

92% EC caused severe irritation in the Draize assay; 91.3%, 100% technical products, 0.5%–0.67% mixed RTU formulations caused minimal irritation

RTU mixed products 0 with 25% and 0.49% PBO were nonsensitizers in the Buehler assay

Listed in registry data as a component of pyrethrin mixture mixtures.

n-octylbicycloheptenedicarboximide

113-48-4

Two mixed products containinng 4.8% and 13.4% NOBD caused minimal irritation in the Draize assay. A product with 12.5% NOBD caused moderate irritation.

Studies with mixtures only

0

Listed in registry data only as a component of mixtures.

40

15 cases involving Type I and 25 case involving Type II pyrethroids, discussed below; there were an additional 66 cases that involved mixtures with synergists that did not meet the case definition for the “pesticide handler” database

Pyrethrins/pyrethroids

Hayes’ Handbook of Pesticide Toxicology

Synthetic pyrethroids

Table 28.6  (Continued) Compound or group Identifiers

Data from pesticide handler data base 1982–2006

CAS #

Draize irritation test

Sensitivity

# of cases Case examples

allethrin

584-79-2, 42534-612, (cis/trans allethrin)

A cis-trans technical d-allethrin, caused corrosion (category I irritation). Other technical formulations and RTU caused minimal irritation

Nonsensitizer

0

1994–138: A worker spilled a mixture of piperonyl butoxide and allethrin on the outside of a backpack sprayer and his arms while mixing a tank load and wiped off the sprayer with a paper towel. After applying the material, he noticed itching, redness, and swelling of the lower back and arms. (case not included in in handler data base because it involved a mixture).

bifenthrin

82657-04-3 13 products (0.18%AI–94% AI) caused minimal irritation in the Draize assay. 3 products (2.4%AI–13.2% AI) caused moderate irritation and 3 caused severe irritation (13%AI–24.9% AI).

14 products 0.184% AI–88.3% AI were negative in the Buehler test. 2 products (4% EC and 26% liquid) were both positive in same assay.

2

1994–1052: An applicator bumped his right arm against a spray nozzle and got some of the bifenthrin spray solution on the arm and then developed numbness and tingling in the right arm. 1995–1210: A mixer/loader, employed by a professional agricultural pest control company to treat cotton, splashed bifenthrin on his arms, face and eyes while transferring product from a closed system holding tank into a 1-gallon container. The container overfilled and the pressure created forced the product out. He develped redness, and a burning sensation on the face, chest, and shoulder.

permethrin

52645-53-1 9 products (0.5%–36.8% EC) caused moderate-severe irritation

3 products (25%–95.6% 8 AI) positive in GPMT; 4 products (0.2%–30.2% AI) positive in Buehler test. A 45% liquid caused sensitization in the LLNA.

92–1381: Worker was mixing material and small amounts kept getting under gloves and shirt. He developed pain, swelling, and blisters on hands and forearms.

phenothrin

26002-80-2 10% formulation of phenothrin and 6 end use mixtures (with NOBD, isopropanol, quaternary ammonium compounds and tetramethrin); 3 products (0.1%–0.4% AI), caused moderate irritation

RTU mixture with tetramethrin equivocal sensitizer in the Buehler test

2

89–1960: A restaurant worker made application of an aerosol pesticide and the spray contacted his left arm and hand. The areas exposed to the spray began to swell and turn red.

resmethrin

10453-86-8 85% technical material is category III irritant; dilute materials caused minimal irritation

A product containing 3% resmethrin aqueous concentrate caused sensitization in the Buehler assay. Six products (0.05%–4% AI) caused no sensitization.

3

2000–122: As a 41-year-old man applied an aerosol insecticide, some of the product contacted his lip. His lip became irritated so he rinsed it with hydrogen peroxide, which worsened the condition before he sought medical attention.

tetramethrin

2117279

RTU mixture with tetramethrin equivocal sensitizer in Buehler study

0

Type I pyrethroids

21% mixed with 21% resmethrin causes transient irritation

715

(Continued )

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

Compound

Predictive Tests in Animals

716

Table 28.6  (Continued) Compound or group Identifiers Compound

Predictive Tests in Animals

CAS #

Draize irritation test

Data from pesticide handler data base 1982–2006

Sensitivity

# of cases Case examples

12

Type II pyrethroids 1999–497: A mixer/loader failed to completely close the shutoff lever of his closed system causing pressure to build up in the connecting hose. When the hose burst, dilute cyfluthrin sprayed onto the unprotected skin on his face. He developed itching, burning and red skin on the right side of the face.

68359-27-5 20% WP and 6 dilute RTU products caused minimal irritation in the Draize assay. A 1% pour-on product for cattle ectoparasites, a 6% AC, and 2 different ECs containing 25% cyfluthrin caused moderate irritation

Mild sensitization response caused by 25% EC; similar 25%, 0.1% granular, and RTU aerosol negative

cyhalothrin

91465-086, 6808585-8

10% cyfluthin/13% PBO 2 in a cattle ear tag, 24% granular caused weak/ equivocal sensitization in 3 formulations. A 9.7% AC a weak sensitizer in GPMT. Two end-use formulations (0.03% and 0.05%) were negative in the Buehler assay.

1998–1191: When an irrigator switched valves on the irrigation pump, water sprayed onto his face and arms. Mildly red, itching and irritated skin on the forearms and face. 2003-497: After mixing cyhalothrin in a 2-gallon sprayer to apply around the warehouse, a worker rubbed his eyes with his bare hands. He developed conjunctivitis and burning skin around the eyes a few minutes later.

cypermethrin

52315-07-5 26% emulsifiable concentrate caused severe irritation. 6 formulations (2 25.3% emulsifiable concentrates, an 18.1% emulsifiable concentrate, and 3 formulations containing less than 1% cypermethrin) caused moderate irritation; 9 products (24.8% EC, 17.1% EC, 16% ear tag, and RTU formulations) caused minimal irritation

0.824% AI in ethylene 5 glycol mild sensitization in Buehler assay; 1% spray negative. Negative in LLNA; positive in GPMT

1988–2388: A mixer/loader handling cypermethrin developed burning in the groin area, shortly after going to the bathroom without thoroughly washing his hands.

deltamethrin

52918-63-5 98% technical and 10 formulations (0.01%–4.95%) caused transient irritation; 11.7% gel and 2.86% EC caused severe irritation

Two dilute formulations (0.01% deltamethrin) were nonsensitizers in the Buehler assay.

1

2001–799: An SPCO applied deltamethrin to the exterior of a restaurant. As he took off the backpack sprayer, the spray wand hit an object that triggered a shot of pesticide directly into his face. He immediately washed up, but developed red, itchy and burning facial skin, “bulging eyes” later that day.

esfenvalerate

66230-04-4 24.6% EC, a 9.53% suspension, a 3.48% concentrate, 2 dilute RTU (1% AI) caused moderate irritation; 0.443% EC caused severe irritation. 7 products (35% WP, 18% liquid, and 5 dilute products (1% AI) caused minimal irritation

Nonsensitizer by Buehler 2 method (18% liquid, 9.53% suspension and 4 dilute products (1% AI))

2000–534: While loading a spray tank, a grower knocked the esfenvalerate container against the spray tank and splashed the concentrate on his chest. He decontaminated promptly, but had red, itching and burning skin on the chest, slight tingling sensation on the arms and lips when he sought medical attention in the evening.

Table 28.6  (Continued)

25% ME, 12.7% EC, 11.4% ME, 9.7% AC, 0.05% RTU aerosol formulation, 0.04% granules, and a 0.03% RTU aerosol caused minimal irritation in the Draize assay. 9.53% ME and 2 10% WPs caused moderate irritation

Hayes’ Handbook of Pesticide Toxicology

cyfluthrin

Compound or group Identifiers

Predictive Tests in Animals

Draize irritation test

Data from pesticide handler data base 1982–2006

CAS #

Sensitivity

fenvalerate

51630-58-1 10.5% WD and a dilute mixture (1% Referred to esfenvalerate 1 AI, pyrethrins, PBO and NOBD ) caused studies; no studies on minimal irritation; a dilute mixtures (0.5% fenvalerate fenvalerate, PBO, NOBD) caused severe irritation; another (0.4% fenvalerate, chlorpyrifos,DDVP) caused moderate irritation

Organochlorines

# of cases Case examples 1986–1191: Measuring cup with concentrated fenvalerate in it was about to fall, and some spilled on his arm. He did not wash or change clothes. He developed a burn at the site of contact.

10

dicofol

115-32-2

50% sp, 42% EC caused minimal irritation 50% sp sensitizer in the 4 in the Draize assay. Separate study of the Buehler assay; A Buehler 42% EC showed moderate irritation. study of the 42% EC showed no sensitization.

1984–954: An employee was mixing dicofol for an aerial application on corn, wearing gloves and face shield, when some material splashed up on his neck. He saw doctor 3 days later when the burning and itching on the front of neck did not improve after initial treatment with first aid ointment. The condition was recorded as a second degree burn. 1984–1454: A mixer/loader/applicator splashed dicofol on the arms that soaked through his protective clothing. Erythema of the forearm was noted when he sought treatment 3 days later.

dienochlor

2227-17-0

No data

No data

3

No cases of contact dermatitis following direct accidental exposure

endosulfan

115-29-7

33.7% EC, 50% WP caused minimal irritation in Draize assay

33.7% EC sensitizer in GPMT

0

lindane

58-89-9

No dermal irritation data available for review

20% liquid formulation for control of leaf borers is a sensitizer in the Buehler assay.

3

methoxychlor

72-43-5

25% formulation labeled as minimal irritant

No data

0

Bacillus thuringiensis & 68038-71-1 Minimal (0.5%–8% formulations) – related endotoxins moderate (3.2–6.4% formulations) irritation in the Draize assay Biological insecticides Azadiracthin

0.436% formulation 1 negative in Buehler assay

1986–309: A hose split during an application under a house and the material sprayed onto the applicator’s hands. He made repairs without gloves, washed off, but noticed a rash later in the day.

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

Compound

No cases of contact dermatitis following direct accidental exposure

7 1141-17-6

40% technical material and 7 other formulations (containing 0.15%–4.5% AI) caused minimal (category IV) irritation and two 3% formulations caused moderation (category III) irritation in the Draize assay

Buehler sensitization 0 studies of four formulations (0.15%– 1.9% AI) were negative. A guinea pig maximization study of a 1.8% EC was positive for sensitization.

717

(Continued )

Compound or group Identifiers Compound

718

Table 28.6  (Continued) Predictive Tests in Animals

CAS #

Draize irritation test

abamectin – mixture of 71751-41-2 9 abamectin products (0.01%–46.3% avermectin Bla and Bib AI) caused minimal irritation. 5 products (0.01–2.21% AI) caused moderate irritation

Miscellaneous insecticides

Data from pesticide handler data base 1982–2006

Sensitivity

# of cases Case examples

Negative in the Buehler 7 test (6 products:0.011%– 49.6% AI); 1.8% EC positive in GPMT

1992–520: An employee developed skin problem on arm after spraying roses with abamectin. He was wearing a rubber rainsuit, but felt wetness on his arm, and did not wash the affected area immediately. Examination showed ulcerative lesions with mild surrounding erythema on right proximal forearm. 1992–2243: A worker developed a rash on his right hand while applying abamectin to roses. He developed a similar rash the previous year after spraying the same pesticide.1998–349 A worker tripped over a wire and accidentally sprayed himself with abamectin on the chin. He developed redness, itching, and burning in the exposed area.

111 33089-61-1 5 EC formulations of amitraz (12.5%– 20% AI) caused moderate (category III) irritation in the Draize assay. A flea collar with 9% amitraz caused only mild irritation.

5% WP caused 0 sensitization in the guinea pig maximization test.

borates

1303-96-4

3 formulations (technical product with 100% AI, & liquids containing 5% and 5.4% AI) caused minimal irritation in the Draize assay

Nonsensitizer in Buehler 2 assay

butoxy polypropylene glycol

9003-13-8

10% butoxypolpypropylene glycol, mixed with permethrin, piperonyl butoxide and pyrethrins caused minimal irritation in the Draize assay

Same mixture containing 0 10% AI caused sensitization in the Buehler assay

Diethyl toluamide (DEET)

134-62-3

22 productsd (7–7% AI) caused minimal (category IV) irritation in the Draize assay. Four formulations (25–25% AI) were nonsensitizers in the Buehler assay.

Contact urticaria 3* reported in public domain literature; Buehler studies negative (30%–40% AI)

hydramethylnon (aminohydrazine) bait

67485-29-4 5 products (0.365%–98% AI) caused minimal irritation in the Draize assay

92% technical product 0 a nonsensitizer in the Buehler assay, a mixed formulation with 0.365% hydramethylnon, 0.25% methoprene was a weak sensitizer

1982–1871: Worker had an allergic reaction (hives) after treating himself with an insect repellent according to the label directions. 93–1422: Worker sprayed an insect repellent on her exposed skin before collecting a lab sample from treated sewage water. She suffered an apparent allergic reaction to the repellent a short time later.

Hayes’ Handbook of Pesticide Toxicology

amitraz

Table 28.6  (Continued) Compound or group Identifiers

Predictive Tests in Animals

Data from pesticide handler data base 1982–2006

CAS #

Draize irritation test

Sensitivity

# of cases Case examples

Imidacloprid

10582778-9

Minimal irritant in 31 formulations tested, Nonsensitizer in the 1 3 showed moderate irritation Buehler assay and GPMT

k salts of fatty acids (insecticidal soaps)

61790-44-1 Some concentrated products 19–50% AI cause moderate-severe irritation or corrosion in the Draize assay

Nonsensitizer in the Buehler assay

1

2005–223: an applicator got spray on her arms while applying insecticidal soap while wearing a sleeveless shirt and developed a corresponding rash

oxythioquiuox

2439-01-02 3 products (25%–92.3% AI) caused minimal irritation in the Draize assay.

40% flowable concentrate is a weak sensitizer in the Buehler test

2

1982–869: While filling a nurse tank, some of the material overflowed, landing on the mixer/loader’s face, neck and arms. Subsequent clinical examination showed marked irritation of the skin in the exposed areas.

propargite

2312-35-8

Technical material (listed as 90.6% AI) 30% wettable powder and the liquid formulation used on cotton nonsensitizer in Buehler (73.86% AI) caused corrosion in the assay Draize assay. The emulsifiable concentrate (69.62% AI) caused severe irritation. Two powdered formulation (28.99% AI and 32% AI) nevertheless caused minimal irritation in the Draize assay.

105

1985–1909: ground applicator contaminated his shirt with his hands – developed burns on his chest. 1982–1667: Splashed a few drops of concentrate on his neck in opening a can for closed system loading. Developed a rash, which persisted for a week until he got it treated.

sulfuramid (bait)

4151-50-2

3 products (0.5%–99% AI) caused minimal irritation in Draize assay

Buehler study negative 1 for mixture of sulfuramid and chlorpyrifos

2002–940: A landscape employee helped on a pesticide application by driving the truck. He helped on landscape maintenance in between sites. At one site, the pesticide dripped from a tree onto the back of his neck where he developed a red and slightly ulcerated rash the next day.

2001–185: An apartment complex employee placed termite bait stations without wearing hand protection. Some of the termiticide contacted his forearms. He developed a rash on the exposed area.

Fungicides Phthalimido compounds

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

Compound

15

captafol

2939-80-2

Labeled as irritant, but no data available for review

captan

133-06-2

folpet

133-07-3

Sensitizer per public domain literature

1

No cases involving dermatitis following direct contact

3 formulations (38.52%–90% AI) minimal Negative in Buehler irritant in Draize assay assay; Sensitizer per public domain literature

14

1996–258: employee was loading captan into the tank of an orchard sprayer when the powder blew up under his face shield. He developed burning, red and itchy skin on the face and neck.

88% technical solid, dilute mixture (0.5% Technical material folpet, 0.5% bis(tributyltin) oxide) caused sensitizer in GPMT minimal skin irritation in the Draize test; 49.4% liquid folpet concentrate caused moderate irritation.

0

No cases in the pesticide handler database

(Continued )

719

720

Table 28.6  (Continued) Compound or group Identifiers

Predictive Tests in Animals

Data from pesticide handler data base 1982–2006

Compound

CAS #

Draize irritation test

Sensitivity

# of cases Case examples

plondrel

5131-24-8

No registration data

Sensitizer per public domain literature

0

Carbamates

14

Benomyl

17804-35-2 Reported as minimal irritant in public domain literature

Sensitizer in the guinea 13 pig maximization test per public domain literature

1985–448: An applicator splashed benomyl onto his face and neck while spraying pruning cuts in vineyard. He suffered burning, erythema, irritation, and swelling of eyes and face. The condition was described as a first degree chemical burn.

Thiophanate methyl

7912735

2 products (an 85% granule, a mixture of 28.5% thiophanate and 51.42% flutolanil) negative in Buehler assay; the 96.2% dust formulation caused sensitization in the GPMT.

No reported episodes of contact dermatitis following direct exposure

11 products (1.5%–96.2% AI) caused minimal irritation in the Draize assay.

Thiocarbamates

1

16 8018-01-07 22.1% copper sulfate, 30.4% mancozeb mixture caused moderate irritation in the Draize assay. 6 products (15%–80% AI) caused minimal irritation.

47.8% liquid, mixture 3 of 63% mancozeb/15% thiophanate methyl caused sensitization in the Buehler assay, 2 products (33.9% liquid, 82.3% powder) negative in the same assay. 30.4% mancozeb/22.1% copper sulfate, mancozeb technical powder sensitizers in GPMT

1986–619: Developed rash on neck after application; has hixtory of sensitivity to mancozeb.

Maneb

12427-38-2 38.8% maneb liquid caused moderate irritation; mixed powder (8% maneb, 0.01% Streptomycin sulfate) caused minimal irritation.

38.8% maneb liquid 2 nonsensitizer in Buehler test; reported as sensitizer in human case reports

1984–811: An employee developed an allergic rash after spraying dithane on grapes

Hayes’ Handbook of Pesticide Toxicology

Mancozeb

Table 28.6  (Continued) Compound or group Identifiers

Predictive Tests in Animals

Data from pesticide handler data base 1982–2006

CAS #

Draize irritation test

Thiram

137-26-8

98.8% AI, 77% granule, 2 dilute mixtures 77% granule sensitizer caused minimal irritation in the Draize in Buehler assay; 1% assay. concentrations showed sensitization in LLNA domain literature

zineb

12122-67-7 No registration studies available for review

Labeled as a sensitizer

ziram

137-30-4

Labeled as a sensitizer

96% industrial formulation labeled as an irritant

Sensitivity

Copper compounds

# of cases Case examples 5

1984–1488: A worker treating seeds with thiram dust developed lesions around the respirator line on the day of the application.

6

1983–298: A worker applied ziram with no hand or face protection and developed contact dermatitis. 1984–518: An applicator spraying almonds with ziram developed a rash after a hose broke on his spray equipment. 1987–203: Loading ziram WP when the tractor driver revved the engine & it blew the material in his face & on his body. 2000–632: As a worker loaded ziram into a nurse rig, the wind blew some of the mixture onto his face. After transferring the mixture from the nurse tank to the air blast sprayer’s tank, he flushed his exposed skin with water. He developed a rash by the next day.

19

Bordeaux mixture

No data

No data

1

85% formulation category III irritant

Labeled as sensitizer

2

copper

7440-50-8

copper ammonium carbonate

33113-08-5 w113123_24.1%liquid_cat4.815

Labeled as sensitizer

0

copper hydroxide

20427-59-2

Labeled as sensitizer; Buehler study on CuOH negative

7

copper naphthenate

1338-02-9

8% liquid moderate irritant; 2%, 8%liquid 80% EC severe irritants; 40% BORAX, 18.16% cu naphthenate corrosive

68% CuNaphthenate 4 Nonsensitizer in Buehler test

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

Compound

1994–561: A worker applied copper hydroxide to walnuts using a high-volume sprayer. His face began to itch and burn a few hours after he finished the application. The affected area was on the unprotected portion of his face. 1990–1368: A worker applying fungicide to nut orchard, wearing all protective gear, got wet from rain blowing in around his hood and down his gloves. He then began itching in areas that had gotten wet. 1987–1724: A wood worker was painting a copper naphthenate wood preservative to the cut end of wood and developed a chemical burn to his arms. 1990–1127: A student employee wearing rubber boots, gloves, goggles, respirator, and ran-suit treated wooden benches with preservative. He accidently rubbed his neck and face while wearing the rubber gloves and developed contact dermatitis. (Continued )

721

722

Table 28.6  (Continued) Compound or group Identifiers

Predictive Tests in Animals

Draize irritation test

Data from pesticide handler data base 1982–2006

Compound

CAS #

cuprous oxide (I), copper oxide (II)

1317-38-0, 85.7% category III irritant, 97% technical Buehler method 0 1317-39-1 material nonirritant nonsensitizer, Animal tests do not show clear evidence of sensitization

copper oxychloride

1332-40-7

copper sulfate

7758-98-7, 99% technical material minimal irritant 7768-98-7

Technical material nonirritant

Sensitivity

# of cases Case examples

Labeled as sensitizer

1

Sensitizer per public domain literature

4

3.5% formulation minimal irritant

Nonsensitizer

0

cupric oxide

1317-39-I

97% technical minimal irritant

Buehler method nonsensitizer

0

Anilazine

101-05-3

No data available for review

Labeled as a sensitizer 4 based on public domain literature

carboxin

5234-68-4

Technical material is nonirritant

Mixture carboxin and other compoundsnonsensitizers

chloroneb

2675-77-6

30% formulation with 3.5% metalaxyl moderate irritant

Negative test on mixture 0

Miscellaneous compounds

0

1992–732: A mixer-loader for an aerial application developed rash on exposed skin areas while dumping wettable anilazine powder into mix tank. At examination, he had a generalized rash on face, neck, and arms thought to be allergic in nature.

Hayes’ Handbook of Pesticide Toxicology

copper triethanolamine 68027-59complex 6,8202759-6

1990–2391: After adding copper sulfate to water, a worker developed a rash on his forearms and itching all over. 1990–2588: While an employee mixed copper sulfate, some powder got inside the glove causing the irritation to his right forearm. The resulting dermatitis was subsequently complicated by an infection. 1993–1839: A worker applied copper sulfate granules to canal water by a piece of equipment he called the “sandblaster” – that air blasts the material on the canal water. After copper sulfate dust landed on his neck and chest, he developed Iarge pruritic, erythematous patches on the neck and chest.

Table 28.6  (Continued) Compound or group Identifiers

Predictive Tests in Animals

Data from pesticide handler data base 1982–2006

CAS #

Draize irritation test

Sensitivity

chlorothalonil

1897-45-6

20.06% chlorothalonil/10.1% 3-Iodo-2propynyl butylcarbamate liquid mixture corrosive; 50% liquid, 14.7% methylene bisthiocyanate/14.5% chlorothalonil severe irritant; other products moderateminimal irritants

Mixed positive, negative 18 findings in Buehler test and GPMT; positive LLNA: EC3-value for chlorothalonil was determined to be 0.002%

fenarimol

60168-88-9 0.78% granules caused minimal irritation No dermal sensitization in the Draize assay studies available for review

3

flusilazole

85509-19-9 60.7% formulation moderate irritant

Buehler test on 20% formulation negative

0

fosetyl-al

39148-24-8 80% powder, 70.2% granular product caused minimal irritation

80% granular product nonsensitizer in Buehler test

0

imazalil

3554-44-0

31% FC severe irritant; Technical material Technical solid (98.5% AI) minimal irritant (98.5%AI) negative in GPMT. Buehler test on 13.5% formulation negative; patch positive contact dermatitis banana production in Central America, veterinary use Europe

0

iprodione

36734-19-7 75% granules moderate irritant; 3 liquids (23%AI, 41.6% AI , 19%iprodione20.4%thiophanate-methyl), 50% granular product caused minimal irritation

41.6% liquid, mixture with thiophanate, 75% granules nonsensitizers in Buehler test

# of cases Case examples

5

1985–75: Mixer/loader-applicator splashed himself with chlorothalonil while mixing. Developed a rash on his arm and a few spots on his face despite promptly washing the exposed areas. 2002–599: as this foreman walked alongside a tractor during an on-going pesticide application, the breeze picked up and drifted chlorothalonil onto him. He immediately washed off, but developed a rash a few hours later. 2004–730: As an applicator checked the electrical pump of the pesticide injection system, the hose carrying the pesticide burst & splashed dilute pesticide onto his face. He immediately washed off, but developed burning and red facial skin around the safety glasses an hour later. No case associated with accidental direct exposure

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

Compound

1995–262: An employee dipping roses into an iprodione solution began to develop a rash and experience swelling. On examination, he had a macular rash on the arms and legs, slight edema of the feet. 1994–515: Worker had spray mist contact his face while applying 50% iprodione WP to strawberries with a boom sprayer. He developed redness and irritation of the exposed area, described as a chemical burn by the treating physician. 2006–558: A nursery worker spraying iprodione felt the pressure in her spray hose drop, then surge, spraying her in the face. Despite wiping her face with a paper towel, overnight she developed itching, irritation, redness and burning on the cheeks and forehead. On examination, the doctor noted reddened skin on the face.

723

(Continued )

724

Table 28.6  (Continued) Compound or group Identifiers

Predictive Tests in Animals

Data from pesticide handler data base 1982–2006

Compound

CAS #

Draize irritation test

Sensitivity

# of cases Case examples

methylene bis(thiocyanate)

6317-18-6

10% MBTC liquid corrosive; similar 10% product moderate irritant; dilute RTU mixture minimal irritant

Dilute RTU mixture non- 0 sensitizer in Buehler test sensitizer

myclobutanil

88671-89-0 Products containing 25% and 27% emulsifiable concentrates caused severe irritation in the Draize assay. A 19.7% liquid caused moderate irritation. Three liquid products (1.55%–84.5% AI) and two solid products (0.39% powder and 0.62% granules) caused minimal irritation.

Two products (1.55% myclobutanil, 2.5% permethrin concentrate and 6% liquid) caused no sensitization in the Buehler test. A 20% liquid was negative in the guinea pig maximization

4

1997–915: an applicator contacted treated foliage with his left arm while checking a malfunctioning arm on a spray rig; the skin in the contaminated area became irritated that night. Sought medical treatment after 4 days because of itching, and noted to have pruritic, red rash on the left elbow and forearm. 92–1379: A worker overfilled a spray tank and spilled pesticide solution (myclobutanil  adjuvant) on his feet. He rinsed his feet and shoes off with water, but did not remove his shoes. Examination showed pruritic dermatitis, scaling, and crusting of the bottoms of both feet.

pentachloronitrobenzene (PCNB)

82-68-8

A 24.3% liquid severe irritant, 15% granular product moderate irritant; 40% liquid, 2 solid products (95% technical, 25% PCNB / 6.25% metalaxyl dust) caused minimal irritation.

15% granular, 23.8% EC, 1 95% technical material negative in the Buehler assay; possible sensitizer per public domain literature

1990–2181: A worker applying PCNB, wrapped a hose around his waist so he could pull the hose while applying with wand, although he was wearing PPE, he developed a rash in areas in contact with the hose on the wrist, waist and hands

sulfur

7704-34-9

80% sulfur/1.56% imidacloprid, 10% 5 products (10%–80% sulfur/0.25% pyrethrins moderate irritants; AI) nonsensitizers in 17 products (0.2%–99% AI) cause Buehler test minimal irritation

68

Hayes’ Handbook of Pesticide Toxicology

1986–968: while loading sulfur in duster, he got some on himself. He was perspiring during the day, possibly aggravating exposure. 1991–1453: A worker developed a rash after applying sulfur dust with a backpack duster. The backpack duster was old, the canvas was ripped and a substantial amount of the sulfur dust leaked out. He developed a severe rash on both arms and the back of the hands, mild rash on the face. 1997–1117: A mixer/loader stood on a biplane to control the loading of sulfur dust. The dust spontaneously caught fire and caused first and second degree burns to the face and wrists. 2000–485, 486: As 2 employees loaded sulfur dust into an aircraft’s hopper, the sulfur ignited and burned both employees. They suffered first and second degree burns (485: the left ear, arms and hands; 486: on both hands) and were taken immediately to the hospital.

Table 28.6  (Continued) Compound or group Identifiers

Predictive Tests in Animals

Data from pesticide handler data base 1982–2006

CAS #

Draize irritation test

Sensitivity

# of cases Case examples

sulfur dioxide

9/5/7446

No data

No data

14

TCMTB (thiocyanomethylthiobenzothiazole)

21564-17-0 30% formulation corrosive

80% AI caused delayed contact hypersensitivity in guinea pigs

0

Triadimefon

43121-43-3 50% wp, 94.6% technical powder, 41.3%liquid, 50% WP, 25% WP, 1% liquid minimal irritants

94.6% technical powder 4 sensitizer in the Buehler test

vinclozolin

50471-44-8 41.3% liquid, 50% WP minimal irritants

50% WP sensitizer in GPMT

0

thiabendazole

148-79-8

Transient irritation with 98.5% formulation

Tests on mixtures only

0

triadimefon

431217343-3

50% formulation causes transient irritation 95% technical material is a sensitizer in the Buehler test

vinclozolin

50471-44-8 Transient irritation

Labeled as sensitizer

0

542-75-6

Sensitization in applicators reported in the public domain literature

18

4

1998–1184: While a worker prepared to treat table grapes with sulfur dioxide, a valve leaked the material against his thighs. He had red, burning, itchy and dry skin on both legs at the site of contact. 2000–40: While adding sulfur dioxide to grapes, a winery worker apparently spilled some of the liquid on her bare hand. Her hand began itching the next day and blistered a few days later. 1993–1694: A delivery hose came loose during treatment of grapes with sulfur dioxide and hit worker in the face. This resulted in erythema on the right side of the face and a foreign body (with rust ring) in the right eye.

1982–1412: Wind blew spray back on applicator who developed a rash reported as a possible allergic reaction to triadimefon.

82–1412: While applying Bayleton° to grapes, the wind blew spray back on to the applicator. He suffered a reported allergic reaction.

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

Compound

Fumigants 1,3 dichloropropene, d-d mixture

60.3% 1,3 dichloropropene/ 3.2% chloropicrin; 81.2%1,3 dichloropropene/16.5% chloropicrin corrosive; 92% liquid AI without chloropicrin minimal irritant

1982–1055: A worker repairing an injection pump had fumigant spill onto his leg when a hose got disconnected. The area of contact on his left calf developed a third degree burn. 1985–59: A worker sat on a fumigant-contaminated glove and developed a red, ulcerated area on the right buttock. 1983–737: A worker wore leather gloves while cleaning a filter on a fumigant, but developed burns on right arm and hand. 1988–2091: An employee unclogging an injector tube on a rig while wearing leather boots had fumigant spill on his foot. He continued to work but the next day he had blistering and swelling of the foot. (Continued )

725

726

Table 28.6  (Continued) Compound or group Identifiers

Predictive Tests in Animals

Data from pesticide handler data base 1982–2006

CAS #

Draize irritation test

Sensitivity

# of cases Case examples

ethylene dibromide

106-93-4

Severe skin irritant per public domain literature

No data available for review

4

1982–171: EDB dripped from the shanks onto the boot w/o the employee knowing, causing first and second degree burns. 1982–536: An employee checking a clogged injection line suffered a burn on the back of his left leg. 1982–748: A worker suffered a burn when EDB dripped on to his foot while he was mounting a tank on a tractor. 1983–310: priority case, 16-KER83. A loose hose allowed EDB to escape, contacting his hand. He was hospitalized to guard against post exposure reaction.

methyl bromide

74-83-9

Corrosive

No data

76

1987–2721: A worker was using an injection probe to apply methyl bromide to tree holes & dripped some on his leather boot, which he wore for at least 24 hours after exposure. His left foot became severely burned. 1988–299: A worker was fumigating tree holes in an orchard. When he tried to stop ground lead he got some methyl bromide on foot. He continued to work without decontaminating exposed area. He developed 1st and 2nd degree burns. 1992–682: Liquid methyl bromide apparently dripped onto the shoes of a worker who was fumigating tree holes in hard soil. He developed blisters on top of both feet.

methyl iodide

74-88-4

99.7% technical liquid, 98.% CH3I, 2.19% chloropicrin severe irritant; 98% CH3I, 2% chloropicrin corrosive

99.7% technical liquid 0 GPMT positive; 98% AI, 2% chloropicrin Buehler positive; 25% CH3I, chloropicrin 75% Buehler positive

propargyl bromide

106-96-7

Possible methyl bromide replacement, irritant per public domain literature

No data

0

aluminum phosphide

20859-73-S No data

No data

6

1984–2184: A rash developed on torso after application of Phostoxin® under tarp for rice in warehouse. There was history of direct exposure.

ethylene oxide

75-21-8

No data required as minimal dermal contact expected

Some products labeled as sensitizers

5

1987–2720: A hospital worker stuck her hand in a sterilizer, before it had aerated to get rid of the ethylene oxide and suffered a chemical burn on her hand.

dazomet

533-74-4

24% liquid corrosive; 20% liquid products, 98.5% solid reported minimal irritants

Sensitizer per public domain literature

2

1990–2448: Applying dusty granular form of pesticide, a worker developed a rash at the belt-line as well as front of legs, abdomen, and arms.

Hayes’ Handbook of Pesticide Toxicology

Compound

Table 28.6  (Continued) Compound or group Identifiers

Predictive Tests in Animals

Data from pesticide handler data base 1982–2006

CAS #

Draize irritation test

Sensitivity

# of cases Case examples

metam-sodium

137-42-8

Five liquid products (32.7%–43.8% metam-sodium) corrosive in the Draize test; unexpected minimal irritation reported for 3 similar products (32.7%– 42.2% metam-sodium)

42.1% liquid apparent sensitizer, but also markedly irritant; public domain report  reactions to 0.05% metam-sodium in water

59

2004–1134: A worker suffered chemical burns on his forearms and hands while performing metam-sodium treatments. He reported getting the liquid inside his oversized wrist-length gloves during the application. Red and swollen skin with papules on the hands and forearms. 1987–1790: A landscape worker was applying chemical to turf without wearing protective boots, when he got some on his foot; it caused a chemical burn. 1991–2296, priority case 58-IMP-91: After loading metam-sodium, a worker would disconnect the hose without turning off the shut-off valve. He apparently spilled small amounts of the metam-sodium on his leather boots each time he loaded. A severe rash and cellulites developed on his feet that spread to the legs and stomach.

diquat

85-00-7

2.3% diquat with oxyfluorfen, dicamba and fluazifop-p-butyl severe irritant; 3 liquid products (2.3%–37.3% AI) moderate irritants; 4 products (0.23%– 8.35% AI) minimal irritants

5 products (2.3%– 22 37.2% AI) all nonsensitizers in Buehler test

1986–1498: An applicator wet his shoes with diquat and did not change them. He developed a rash on the top of one foot. By the time he saw a doctor, 8 days later, his foot had become infected. 1991–2215 & 91–2552: Two park maintenance workers were sprayed in the face with diquat when the exhaust muffler on the engine burned a hole through the spray hose. They washed their skin immediately, but still developed some redness on the face.

paraquat

1910-42-5, Irritant 2074-50-2, 4685-14-7 depending upon salt

22.3% and 37.1% AI 35 nonsensitizers in Buehler test

83–480: A gust of wind blew material onto arms which had previous abrasions and his condition was aggravated by contact with paraquat.

acetochlor

34256-82-1 No data

No data

alachlor

15972-60-8 42.2% EC caused minimal irritation; 45.1% EC and 92.8% showed borderline minimal vs. moderate irrtation (technically inadequate studies)

69.6% granular 1 formulation positive in the Buehler assay; public domain case report of allergy after accidental direct exposure

allidochlor

93-71-0

Public domain case 0 report of allergy after accidental exposure from spill on the feet

Herbicides Bipiridyls

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

Compound

Chloracetanilides

No data

0 1984–537: An applicator developed a fine rash on trunk, arms, and legs two years in a row after handling alachlor. His condition was reported as suspected allergic dermatitis.

727

(Continued )

728

Table 28.6  (Continued) Compound or group Identifiers

Predictive Tests in Animals

Draize irritation test

Data from pesticide handler data base 1982–2006

Compound

CAS #

Sensitivity

# of cases Case examples

butachlor

23184-66-9 Technical material moderately irritating

Technical material sensitizer in Buehler assay; case report possible IgE mediated case of butachlor hepatitis

0

metolachlor

51218-45-2 Mixture 22% metolachlor, 22% cyanazine 97.3% technical liquid 1 caused moderate-severe irritation; 85.1% sensitizer in the Buehler EC caused moderate irritation; 84.1% and assay; 84.1% EC negative 97.3% liquid caused minimal irritation category 3 irritant

propachlor

1918-16-7

Reported as severe irritant per public domain literature

Multiple cases of allergic 0 contact dermatitis in bulb growers, Netherlands, 1993

benfluralin (benefin)

1861-40-1

Technical liquid (95.64% AI) moderate irritant; 96.6% solid severe irritant

No data available for review

ethalfluralin

55283-68-6 18.5% liquid mixture with 5.7% clomazone caused minimal irritation

oryzalin

19044-88-3 96.55% powder, 41% liquid, 40.4% liquid, mixture of 1% oryzalin and 1% benefin caused minimal irritation

pendimethalin

40487-42-1 37.4% liquid, 38.7% suspension minimal 38.7% liquid non2 irritants sensitizer in Buehler test

No cases of dermatitis following direct contact

trifiuralin

578064

1993–340: An applicator was loading his tractor with trifluralin and some of the material leaked out and was blown onto his face. He failed to wash the exposed area right away and developed developed an itching, burning, red rash on face. 1994–565: Herbicide sprayed applicator in the face and caused facial dermatitis

Nitroaniline compounds 1988–29: Wind blew the material onto him while he was pouring benfluralin (Balan) into a mix tank. Developed rash & itching in areas on contact on trunk & extremities.

1

1990–1832: An applicator disconnected a filter valve, was sprayed in the face with herbicide, and developed pruritus.

40.4% liquid non7 sensitizer in Buehler test

2% trifluralin, 0.25% isoxaben, 0.25% oxyfluorfen mixture sensitizer in GPMT; 43% liquid negative in Buehler test

14

1984–51: The wind blew material into an applicator’s face, resulting in a rash on neck and face. 1984–272: An applicator wiped his face with a wet glove, and developed a rash immediately.

Hayes’ Handbook of Pesticide Toxicology

Organophosphate-like compounds

A 50.8% EC caused moderate irritationy. Seven products (0.74% granules – 80% dry flowable formulation) caused minimal irritation.

1

Table 28.6  (Continued) Compound or group Identifiers

Predictive Tests in Animals

Data from pesticide handler data base 1982–2006

CAS #

Draize irritation test

Sensitivity

# of cases Case examples

bensulide

741-58-2

Solid and liquid products, 3.6%–92.5% AI, minimal irritants in the Draize assay

Inadequate

2

1983–1770: A hose ruptured and sprayed his arm. A rash developed, reported to cause 20 days lost-work-time.

glyphosate, sulfosate, and derviative salts

1071-83-6, 5 liquid products (28.3%–48.8% AI) 40.1% liquid sensitizer 81591-81-3 moderate-severe irritants; 9 liquid, 1 solid in LLNA, estimated EC3 (1-62%AI) caused minimal irritation 7–8% of formulated product; all products negative in Buehler test

204

1983–220: The wind blew spray mist onto a worker’s forehead while he was applying a glyphosate formulation. He experienced a rash and itching at the site of contact that lasted for several days. 1983–917: As a worker was treating vineyard weeds with a glyphosate formulation, a hose burst on his backpack sprayer, covering his back with the material and he subsequently developed a rash at the site of contact.

tribufos

78-48-8

71% liquid corrosive in Draize assay; 99.7% liquid caused moderate irritation.

0

2,4-D

94-75-7

2 powder, 6 liquid products (0.2%–96.7% 19.64% liquid sensitizer 7 AI) minimal irritants; 3 mixtures caused in the Buehler test; other moderate irritation 4 products negative; India case series 3  patch tests

dicamba

1918-00-9

86.8% solid technical material minimal irritant

86.8% solid technical material sensitizer in Buehler test

MCPA

94-74-6

Liquid product containing 51.9% MCPP, granules containing low concentrations of MCPA (0.82%), MCPP (0.33%) minimal irritants

DMA salts of MCPA 1 (40.42%), MCPP (15.97%), and dicamba (3.97%); liquid product with DMA salts of MCPP (10%) and MCPA (14%) nonsensitizers (Buehler assay)

MCPP

7085-19-01 51.9% liquid minimal irritant

See above data for MCPA 0

dithiopyr

97886-45-8 22.9%EC caused severe irritation; 13.5% liquid, 41.4%–91.5% solid caused minimal irritation

12.7% EC sensitizer in Buehler test, 22.9% EC LLNA positive

imazethapyr

81335-77-5 97% aqueous paste (technical 22.9% formulation 0 imazethaphyr), 70% granules, and 21.6% negative in Buehler study aqueous concentrate caused minimal irritation in the Draize test

No data

Phenoxy herbicides

0

1989–15: Contact dermatitis developed on arm after he slipped and spilled herbicide containing MCPA on the affected area

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

Compound

Pyridine derivatives 0

(Continued )

729

730

Table 28.6  (Continued) Compound or group Identifiers

Predictive Tests in Animals

Compound

CAS #

triclopyr

Draize irritation test

Data from pesticide handler data base 1982–2006

Sensitivity

# of cases Case examples

5721-4069- 61.6% formulation causes transient 1 irritation in the Draize test

Buehler study on 17% formulation was negative. However, the 44% formulation is labeled as sensitizer

3

88–2832: The worker was spraying weeds along the roadside when a big rig drove by causing a shift in the wind direction and the spray blew back in her face. She developed mild erythema on the face.

molinate

2212-67-1

15% granules a nonsensitizer in the Buehler assay

1

87–937: An employee was loading molinate bags into the bucket of the loader truck (for an aerial application). Some of the molinate got inside his protective clothing, contacting his legs and feet. He subsequently developed a rash in the corresponding areas.

thiobencarb

28249-77-6 15% granule, 84% EC, 97.4% liquid caused minimal irritation

15% granules a nonsensitizer in the Buehler assay

0

Carbamates 10.3% and 15.2% granular products caused minimal irritation in the Draize assay

Triazines 1912-24-9

1.16% granules, 40.8% liquid, 33.1% 40.8% liquid mixture with 26.1% metolachlor minimal nonsensitizer in Buehler irritants test

cyanazine

21725-46-2 97.3% solid caused minimal irritation in the Draize test caused minimal erythema in the Draize test

Buehler study for EC mixture of 22% cyanzine, 22% metolachlor negative

0

prometon

1610-18-0

97% technical solid, a 45.3% liquid, and mixtures of 2,4-D/ prometon and diquat dibromide/prometon nonsensitizers in the Buehler assay

1

A 1.86% liquid product caused severe irritation in the Draize assay. 2 liquids (2% and 12.5% AI), 3 mixed RTU products (3.59% prometon/1% 2,4-D, inactive mixture 2.5% prometon/1% pentachlorophenol) caused moderate irritation. 3.75% RTU liquid, 45.3% EC, 97.3% technical solid caused minimal irritation.

0

1991–923: Dermatitis following chemical exposure to the face

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atrazine

Table 28.6  (Continued) Predictive Tests in Animals

Data from pesticide handler data base 1982–2006

Compound

CAS #

Draize irritation test

Sensitivity

# of cases Case examples

prometryn

7287-19-6

97.3% technical solid caused minimal irritation in the Draize assay

45% material minimal reaction in Buehler test

0

simazine

122-34-9

90% formulation minimal irritant

6.3% liquid simazine negative in GPMT

1

bromacil

314-40-9

3 DF products (53% bromacil, 27% diuron; 40% bromacil, 40% diuron; 80% bromacil), dilute mixture of 1.5% bromacil with sodium chlorate and sodium metaborate caused minimal irritation

40% bromacil, 40% 0 diuron DF nonsensitizer, Buehler test

chlorsulfuron

6490272-3

75% DF minimal irritant

No data

Diuron

330-54-1

81% formulation category III irritant

Nonsensitizer in Buehler 0 test

halosulfuron

13539730-7

4 powdered or granular products (12.5%– Granular mixture with 1 98.5% AI) minimal irritants dicamba nonsensitizer in the Buehler test

rimsulfuron

12293148-0

Products (25% powder, 25% soluble granules, and 98% technical solid) minimal irritants

sulfometuron methyl

thidiazuron

83–2394: Mix/loading material, apparently urinated during operation, depositing material on penis and he developed a secondary rash.

Urea herbicides

0

25% granular product negative in LLNA

1

74223-56-6 75% granular, DF products caused minimal irritation

75% granular product nonsensitizer in Draize assay

1

51707-55-2 12% thidiazuron/ 6% diuron mixture minimal irritant

12% thidiazuron/ 6% diuron mixture nonsensitizer in the Buehler test.

0

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

Compound or group Identifiers

Miscellaneous (Continued )

731

732

Table 28.6  (Continued) Compound or group Identifiers

Predictive Tests in Animals

Data from pesticide handler data base 1982–2006

Compound

CAS #

Draize irritation test

Sensitivity

# of cases Case examples

bromoxynil

1689-84-5

EC products containing 55% and 33.4% bromoxynil caused minimal irritation in the Draize assay.

33.4% EC sensitizer in the Buehler assay.

2

dichlobenil

1194-65-6

98.8% technical solid, a 15.12% liquid, and a 0.55% dust minimal irritants

15.3% AC sensitizer in GPMT

0

endothall

145-73-3

fluazifop-butyl

69806-50-4 24.9% liquid product caused moderate irritation; dilute RTU mixture with 2,4-D and diquat caused minimal irritation.

No data available for review

flumetsulam

98967-40-9 Mixture with metolachlor category 2 irritant

Mixture with metolachlor 0 sensitizer in Buehler test

Iioxaben

82558-50-7 Technical isoxaben (approximately 93% AI), 75% dry flowable caused minimal irritants

75% DF nonsensitizer in 0 the Buehler test

1985–937: employee exposed while applying material to weeds when the wind shifted exposing herself. 2000–132: An applicator noted a rash on his neck after applying bromoxynil to garlic fields for 5 days. He possibly touched his skin with contaminated gloves. The previous week, he applied paraquat to cotton. Report of eyewear with full-face respirator unresolved. Swollen and itchy rash on the back of the neck and face.

1989–1625: Worker spilled a category 1 endothall product on his legs while hand pouring it, and noted immediate pain despite rinsing. He was taken to the doctor the next day and found to have first and second degree burns over the affected area. 2000–624/625: Two employees applied endothall to a 1.5-acre pond, alternating between applying the herbicide and rowing a boat. When the backpack sprayer quit working, they poured the contents into a hand can. Both had itchy, swollen rashes on the back and reported to the treating doctor that the sprayer leaked.

14

1992–2125: Worker blew on nozzle of his backpack sprayer in an attempt to remove debris from the plugged nozzle. He later developed above dry and cracked lips. 1987–1499: A sprayer hose leaked while a worker was applying fluazifop-butyl for roadside weed control. She removed her spraysuit, rolled up her shirt sleeves & washed her arms, but developed itching in the exposed area. 1989–2100: Using backpack sprayer got herbicide on foot and legs. When it became too painful to walk, he went to the doctor and proved to have cellulitis in the affected areas.

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3

Table 28.6  (Continued) Compound or group Identifiers

Predictive Tests in Animals

CAS #

metribuzin

Draize irritation test

Sensitivity

# of cases Case examples

21087-64-9 Two 75% dry flowable products, 41% flowable concentrate minimal irritants

75% DF products, 41% FC negative in Buehler test

1

No cases of direct accidental exposure

MSMA (arsenical)

2163-80-6

18.7% MSMA mixture, 51% MSMA negative in Buehler test

3

1985–1368: While loading a self-propelled sprayer, worker splashed material on self, resulting in neck and arm irritation.

norflurazon

27314-13-2 5% granules, 99.6% solid minimal irritants

Nonsensitizer in LLNA

1

1989–2739: Employee spraying herbicide wearing full protective gear. In the afternoon develops red rash, peeling skin and weeping about face. Similar reaction several years ago when applying same herbicide; considered extremely sensitive to product.

oxadiazon

19666-30-9 50% wettable powder reported corrosive in one test, miminal irritant in another; 1.4% granules moderate irritant; mixture 1% oxadiazon/0.5% prodiamine minimal irritant

4

1988–813: He applied granular herbicide using his bare hands and his hands were sweaty during the application. His hands broke out in a papular rash.

oxyfluorfen

42874-03-3 97.1% solid technical material, 2 liquid products (41%–42.09% AI) and 3 liquid mixtures (1%–21% oxyfluorfen, combined with glyphosate, oxadiazon or oryzalin) minimal irritants

42.09% liquid, 41% AC, 3 23% liquid, a granular mixture of 2% oxyfluofen and 1% oxadiazon nonsensitizers

1994–858: A worker was trying to fix a pump, that was not working properly, when some oxyfluorfen spilled on his left thigh. He wanted to transfer the oxyfluorfen from a drum into a “microjet irrigation system” tank. He failed to wear the required apron. Developed tingling and itching in the affected area of his leg. 1983–63: In the process of cleaning filters, a worker got mist on his arms. A rash developed that lasted for 2 wks.

picloram

1918-02-01 20.4%, 38.8% liquid products minimal irritants

no data

1

No cases following direct accidental exposure

propanil

709-98-8

80% propanil/0.62% bensulfuron granules, 60% DF nonsensitizers in Buehler test

1

No cases following direct accidental exposure

sethoxydim

74051-80-2 13% liquid caused severe irritation

13% formulation nonsensitizer in maximization test

7

1988–1253: A worker was pumping up sprayer when leaky gasket characterized as a chemical burn.

Liquid products with 9.81% MSMA, 18.7% MSMA & dilute phenoxy herbicides moderate irritants

80% dry granule, 60% dry flowable preparation, 41.2% propanil/ 0.32% bensulfuron-methyl liquid minimal irritants

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

Compound

Data from pesticide handler data base 1982–2006

Abbreviations: AC – aqueous concentrate DF – dry flowable EC – emulifiable concentrate; FC – flowable concentrate ME – microencapsulated; WP – wettable powder; LLNA – local lymph node assay; GPMT – guinea pig maximization test * Cases not in handler database because reactions occurred in end-users not coded as pesticide applicators in the California illness registry database.

733

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28.2.1.4  Phenolic Compounds A review of the CDPR label database showed 128 separate phenol compounds with CDPR chemical codes, including sodium phenate salts of several phenol derivatives (CDPR, 2009). Physical Properties  Ortho-phenylphenol physical properties: formula, C12H10O; MW, 170.21; MP, 59°C; BP, 286°C; log P, 3.09; VP solubility in H2O, 700 mg/l; other solubilities: soluble in fixed alkali hydroxide solutions and most organic solvents Irritation data  O-phenylphenol is a typical compound (see Table 28.6). It is commonly mixed with other phenol compounds, and the mixtures can be very irritating at sufficient concentrations (see Table 28.6).

Irritation, sensitization, and illness data  Organotin compounds have been reported as irritants (Gammeltoft, 1978). Registration studies showed that several mixtures of concentrated tributyltin compounds (10–30% concentrations of tributyltin fluoride and tributyltin methacrylate) caused corrosion in the Draize assay. Less concentrated allergenic effects of tributyl tin compounds generally are probably not significant, but products containing tributyltin benzoate and tributyltin methacrylate were both sensitizers in the Buehler test (see Table 28.6). The single case associated with tributyltin compounds in the handler database involved dermatitis following accidental direct contact (1990-278).

28.2.1.6  Other Antimicrobial Compounds O CHBr2

Sensitization data  O-phenylphenol has been reported to cause no sensitization in the GPMT (Andersen, 1984). Nevertheless, its sodium salt has been identified as a cause of the immediate sensitization reaction, contact urticaria (Tuer et al., 1986). p-tert-Butylphenol is notable for being identified as a human sensitizer and as an occasional cause of occupational leukoderma (Mancuso et al., 1996; O’Malley et al., 1988). Reports of cutting oil dermatitis associated with positive patch test reactions to o-phenylphenol have been published by Adams and Manchester (1982) and Van Hecke (1986). Illness data  Fifty-six cases associated with phenol disinfectants were identified in the pesticide handler database. The data are limited, as with the isothiazolin compounds, because the individual phenols involved in each case were not identified. Typical cases occurred following direct contact. For example, case 1988-909, shown in Table 28.6, involved a simple irritant dermatitis in a janitor associated with use of a phenol disinfectant.

28.2.1.5  Organotin Compounds Physical properties Tributyltin oxide physical properties: formula, C24H54OSn2; MW, 596.11; MP, 45°C; BP, 180°C at 2 mm Hg; log P, 4.05; VP, 7.50  106 mm Hg; solubility in H2O, 4 mg/l; other solubilities: miscible with organic solvents Tributyltin methacrylate physical properties: formula, C16H32O2Sn; MW, 374.7; MP, 16°C; BP, 300°C; log P, NA; VP, 2  104 mm Hg at 20°C; solubility in H2O, NA; other solubilities: NA Tributyltin fluoride physical properties: formula, C12H27FSn; MW, 309.034; MP, 260°C; BP, NA; log P, 4.39; VP, 3.52  106 mm Hg at 25°C; solubility in H2O, 6 mg/l; other solubilities: NA Tributyltin benzoate physical properties: VP, 1.5  106 mm at 20°C; other data, NA

C

OH NH

CN

N

2,2-dibromo-3-nitril-propionamide

C

CH2

OH

2-(hydroxymethylamino)-ethanol

Br HO CH2

CH3

CH2

I

Br

CH2 OH

NO2

I

Iodine

Br

CH2

C

CH2

CH2 CN

CN 2-Bromo-2-nitropropane-1,3-diol

1,2-Dibromo-2,4-dicyanobutane

Physical properties 2,2-Dibromo-3-nitrilopropionamide physical properties: formula, C3H2Br2N2O; MW, 242; MP, 124.5°C; BP, °C; log P, 0.82; VP, 9.0  104 mm Hg; solubility in H2O, 15,000 mg/l; other solubilities: solubility in acetone and ethanol 35 and 25 g/100 ml, respectively 2-(Hydroxymethylamino)-ethanol physical properties: formula, C3H9NO2; MW, 91.1091; BP, 240.2°C; log P, 0.82; VP, 89 mm Hg; solubility in H2O, 15,000 mg/L; other properties: hydrolyzes to monoethanolamine and formaldehyde (U.S. EPA, 2006b) Bronopol physical properties: formula, C3H6BrNO4; MW, 199.99; MP, 131.5°C; log P, 0.640; VP, 1.26E-05; solubility in H2O, 2.50 10  05 mg/l; other solubilities: soluble in alcohol, ethyl acetate; slightly soluble in chloroform, acetone, ether, and benzene 1,2-dibromo-2,4- dicyanobutane (bromothalonil) physical properties: formula, C6H6Br2N2; MW, 265.94; MP, 52°C; log P, 1.630; VP, 2.5  104; solubility in H2O, 1300 mg/l; other solubilities: very soluble in dimethyl­ formamide, acetone, chloroform, ethyl acetate, benzene; soluble in methanol, ethanol, and ether Glutaraldehyde physical properties: formula, C5H8O2; MW, 100.13; MP, 14°C; BP, 188°C; log P, 2.49; VP, 0.17 mm Hg at 20°C; solubility in H2O, miscible; other solubilities: miscible in ethanol Iodine physical properties: formula, I2; MW, 253.809; MP, 113.7°C; BP, 184.4°C; log P, 2.49; VP, 2.33  101 mm Hg; solubility in H2O, 330 mg/l; other solubilities: dissolves readily in chloroform, carbon tetrachloride, or carbon disulfide to form purple solutions

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

Irritation, sensitization, and Illness Data  In animal testing, many of the miscellaneous disinfectant compounds [2,2-dibromo-3-nitrilpropionamide, 2-(hydroxymethylamino)-ethanol, 2-bromo-2-nitro-1,3-propanediol (bronopol), and iodine] and glutaraldehyde appear to be corrosive or severely irritant (see Table 28.6). Highlights of registration data and public domain literature on selected compounds are discussed next. Bronopol has long been recognized as a sensitizer in the public domain literature (Camarasa, 1986; Frosch et al., 1990), related to its capacity for releasing formaldehyde (Cronin, 1980; Kranke et al., 1996). 1,2-Dibromo-2, 4-dicyanobutane (Tosti et al., 1995; Vigan et al., 1996) and iodine (Ancona et al., 1985; Erdmann et al., 1999) have also been associated with cases of contact sensitization. Glutaraldehyde is recognized as a sensitizer in the public domain literature (Endo and Takigawa, 2006). Registration data showed that 2,2-dibromo-3-nitrilpropionamide is also a sensitizer in the Buehler test (see Table 28.6). Organotin compounds have been reported as irritants (Gammeltoft, 1978). Registration studies showed that several mixtures of concentrated tributyltin compounds (10–30% concentrations of tributyltin fluoride and tributyltin methacrylate) caused corrosion in the Draize assay. Less concentrated allergenic effects of tributyltin compounds generally are probably not significant, but products containing tributyltin benzoate and tributyltin methacrylate were both sensitizers in the Buehler test (see Table 28.6). Among the miscellaneous antimicrobial compounds, only iodine and glutaraldehyde were associated with cases in the handler database (see Table 28.6).

735

(see Table 28.6). In the public domain literature, some OPs, such as parathion and malathion, have been reported to cause contact sensitization or other skin reactions. Exposure sufficient to cause skin reaction often poses a risk of systemic poisoning (Mathias, 1983). Details of the animal studies and 77 pesticide applicator cases from the handler database are discussed for selected compounds. O P CH3O

28.2.2.1  Organophosphates The organophosphate (OP) compounds are generally thought to cause minimal irritation (Rycroft, 1977). However, a few compounds, such as dichlorvos (DDVP) and naled, have reactive halogen moieties [element IX in the DEREK model and halogenated alkanes and alkenes (group CNS) in the SICRET model] that can cause acute irritation. Many OPs cause sensitization in the GPMT, and some also cause sensitization in the Buehler test or in the LLNA

NH

C

CH3

Acephate

Acephate Physical properties  Formula, C4H10NO3PS; MW, 183.2; MP, 82–89ºC/technical grade, 82–90% purity; VP, 1.7  106 mm Hg at 25ºC; log P, -0.85; solubility in H2O, 818 g/l; other solubilities (g/l at 20ºC): acetone 151, ethanol 100, ethyl acetate 35, benzene 16, and hexane 0.1 Irritation, sensitization, and illness data  Five products containing acephate powder or pellets (1.5–90% AI) caused minimal irritation in the Draize assay. Three products (1.5–97.4% AI) were nonsensitizers in the Buehler assay. Most of the five cases in the handler database were consistent with reversible irritation following direct exposure (e.g., case 86-1084), but symptoms of urticaria in one case (89-2500) suggested possible sensitization. Cl

Cl

28.2.2  Insecticides and Insect Repellants Use of organophosphate and carbamate cholinesterase (ChE) inhibitors has declined since institution of the Food Quality Protection Act in the late 1990s. California agricultural use data, for example, showed 16,207,537 pounds of ChE inhibitors in 1997 and 5,769,785 pounds in 2007. Use of most ChE inhibitors for structural pest control was completely eliminated during the same time period, although malathion and other low- or moderate-toxicity ChE inhibitors are still used for control of garden insects.

O

CH3S

S CH3CH2O CH3CH2O

P

O

N

Cl

Chlorpyrifos

Chlorpyrifos  Chlorpyrifos is no longer used for structural or home pest control, but it is still widely used in agriculture (1,430,034 pounds reported used in California agriculture, decreased from 3,212,165 pounds reported used in 1997). Physical properties  Formula, C9H11Cl3NO3PS; MW, 350.62; MP, 41–42ºC; VP, 1.88  105 mm Hg; log P, 5.0; solubility in H2O, 1.4 mg/l; other solubilities (at 25ºC): isooctane, 79% wt/wt; methanol, 43% wt/wt Irritation, sensitization, and illness data  Technical chlorpyrifos (97.6% AI, liquified prior to application) produced minimal irritation in the Draize assay. Four products (1.33– 45.4% liquid formulations) nevertheless caused severe irritation. Ten products (0.25–42.8% liquid or spray formulations, including 2 mixtures containing pyrethroid insecticides)

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caused moderate irritation. Fifteen products with greater than 1% chlorpyrifos (1.7–75% AI, including 6 granular, powder, or solid formulations and 9 liquids) caused minimal irritation. All 19 products (principally ready-to-use liquid or spray formulations, including several mixtures with pyrethroid or carbamate insecticides) with less than 1% chlorpyrifos caused minimal irritation in the Draize assay. Six products were tested in animal sensitization assays. These included 42.8 and 44.9% ECs that were both negative in the Buehler test. Two formulations (22.8 and 24.66% ECs) caused sensitization in the Buehler assay. A 41% EC was also a sensitizer in the LLNA, as was an EC containing 30% chlorpyrifos and 0.54% -cyhalothrin. From the available information, it is not certain how much of the variation observed in the preceding irritation and sensitization assays was attributable to variation in inert ingredients or attributable to variations in laboratory methods. The 23 cases associated with chlorpyrifos in the pesticide handler database appeared to be instances of mild irritation (see sample case 1985-343, Table 28.6), consistent with the transient irritation occurring in animal tests of technical material. There were no cases reported for which possible sensitization was evaluated by patch testing. O O

S

Cl O

P

OC2H5 OC2H5

Coumaphos

Coumaphos  Powder or dust formulations of coumaphos are used for control of flies and ectoparasites in cattle. California use data showed limited use for 2007. Physical properties  Formula, C14H16ClO5PS; MW, 362.77; MP, 91ºC; log P, 4.13; VP, 9.7  10-8 mm Hg at 20 ºC; solubility in H2O, 1.5 mg/l; other solubilities: slightly soluble in acetone, chloroform, and corn oil Irritation, sensitization, and illness data  Technical coumaphos (98.25% powdered AI) caused minimal irritation in the Draize assay, although the study was graded as technically deficient because of failure to adequately moisten the test material prior to application. Other products tested (a 25% wettable powder, a 1% dust, and ECs with 12.1 and 11.9% AI) also caused minimal irritation in the Draize assay. The 98.25% technical material and 25% wettable powder were both negative in the Buehler assay. An 11.9% EC used for insect control on livestock was a borderline positive

in the Buehler assay: On challenge exposure, erythema and edema were significantly increased compared to the challenge exposures in the same animals. There were no cases associated with coumaphos in the pesticide handler database.

CH3CH2O

P

O

N

N

S

N

S CH3CH2O

CH

CH3

CH3CH2O CH3CH2O

P

O

N

CH3

Diazinon

CH

CH3 CH3

Isodiazinon

Diazinon  California use data for 2007 showed 29,309 applications, for a total of 350,640 pounds applied on orchards, berry crops, row crops, and nursery crops. Physical properties  Formula, C12H21N2O3PS; MW, 304.35; BP, 83–84ºC; log P, 3.81; VP, 9.01  105 mm Hg at 25ºC; solubility in H2O, 40 mg/l at 20ºC; other solubilities: miscible with petroleum ether, alcohol, ether, cyclohexane, benzene, and similar hydrocarbons. Irritation data  Five diazinon products (25% EC, 47.5% EC, 48% EC, 48% AC, and 48% suspension) caused moderate irritation in the Draize assay. The 87% technical material and 19 other products (0.5–87% AI) caused minimal irritation. Sensitization data  Three products (87.5% granular, 47.5% EC, 50% wettable powder, and a flea collar product) caused no sensitization in the Buehler test. A 2% dust was positive for sensitization. However, Matsushita and Aoyama (1981) identified diazinon as a sensitizer in the GPMT. Illness data  The 10 cases in the handler database were consistent with an irritant mechanism (e.g., case 87-2537, Table 28.6). A case report from Australia identified an isomer of diazinon as a cause of porphyria cutanea tarda (Collins et al., 1982). CH3O

O P OCH

CH3O

CCl2

DDVP

Dichlorvos  Dichlorvos is a moderately potent cholinesterase inhibitor that is still used for control of flying insect pests of cattle. California use data for 2007 showed 6376 pounds used on dairies, poultry farms, and other animal facilities and also nursery crops. Physical properties  Formula C4H7Cl2O4P; MW, 221; MP, 25ºC; BP, 234.1ºC; log P, 1.40–2.29; VP, 0.031503 mm Hg; solubility in H2O, 8000 mg/l; other solubilities: slightly soluble in glycerin; miscible with aromatic and chlorinated hydrocarbon solvents and alcohols Irritation, sensitization, and illness data  DDVP is reported as a cause of patch test negative irritant dermatitis

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

in the public domain literature (Breen and Conroy, 1971; Mathias, 1983). The limited Draize irritation data available (for a 1% fly bait containing dichlorvos) showed only minimal irritation. No dermal sensitization study was available for review. The single case in the applicator database appeared to be a straightforward case of irritant dermatitis following direct accidental exposure (1983-1870). CH3O S P CH3O

p­ roduct tested in a different laboratory was reported to cause minimal irritation. A 10% granular product and 75% EC both caused minimal irritation. The 48.4% EC caused sensitization in the maximization test but was negative in the Buehler assay. There were no cases associated with fosthiozate in the handler database. CH3O

S CH2COHNCH3

CH3O

Dimethoate

Dimethoate  Dimethoate is still a broadly used agricultural insecticide. California 2007 use data showed 19,424 applications, for a total of 314,035.1899 pounds applied to orchards, grain, nursery, cotton, row crops, and vineyards and for landscape maintenance. Physical properties  Formula, C5H12NO3PS2; MW, 229.26; MP, 52–52.5ºC; BP, 107ºC; log P, 0.78; VP, 1.875  105 mm Hg; solubility in H2O, 5000 mg/l; other solubilities: soluble in most organic solvents, such as alcohols, ketones, benzene, toluene, chloroform, and dichloromethane Irritation, sensitization, and illness data  Five products (25% wettable powder, ECs containing 43.5–44.7% AI, and the 96% technical formulation of dimethoate) caused minimal irritation in the Draize assay. A 32.7% liquid formulation caused moderate irritation. Buehler sensitization studies on two products (32.7 and 44.7% EC) were negative but were technically limited because of the absence of positive controls. There were three cases in the handler database. In case 83-1880, dermatitis developed after spraying dimethoate, but no direct exposure occurred. The applicator developed symptoms of nausea and headache, suggesting possible systemic OP poisoning. O

O S

737

N

P S O

Fosthiazate

Fosthiazate  Fosthiazate is a nonfumigant nematicide that has limited use for root vegetable crops. There was no reported agricultural use in California in 2007. Physical properties  Formula, C9H18NO3PS2; MW, 283.35; log P, 1.68; MP, 25ºC; VP, 4.20E-06 mm Hg; solubility in H2O, 9850 mg/l Irritation, sensitization, and illness data  A 48.4% EC caused severe irritation in the Draize assay. The same

O

S P

C O2C2H5

S CH

C C OC2H5

Malathion

O

Malathion  Malathion is still widely used as an agricultural insecticide and for garden pest control. 2007 California use data showed 14,037 applications, for a total of 457,974 pounds on orchards, grains, vineyards, row crops, and nurseries and for landscape maintenance. Physical properties  Formula, C10H19O6PS2; MW, 330.4; MP, 2.8ºC; VP, 3.38  106 mm Hg; log P, 2.36; solubility in H2O, 143 mg/l; other solubilities: miscible with alcohols, esters, ketones, ethers, aromatic and alkylated aromatic hydrocarbons, and vegetable oils Irritation 57% EC) Four ECs tation, as material.

data  Two products (a 44% liquid and a caused moderate irritation in the Draize assay. tested (18.57–80.75% AI) caused minimal irridid a 5% dust and the 96.5% liquid technical

Sensitization data  Minimal animal test data were available to assess sensitization: A Buehler test conducted on a 57% formulation was the only study available for review. Challenge applications showed erythema and caused some eschar formation, but the study did not employ adequate negative controls to verify that the dermal reactions observed were due to sensitization rather than irritation. A public domain report by Milby and Epstein (1964) identified malathion as a sensitizer, but the study did not employ currently accepted standards for evaluating irritation threshold of the test material in unexposed control subjects. Illness data  Most of the 10 cases associated with malathion in the handler database were consistent with mild, transient irritation (see sample case 1994-401, Table 28.6, involving an episode of accidental direct exposure). CH3O O P NH2 CH3S Methamidophos

Methamidophos  Methamidophos is a potent cholinesterase inhibitor. California use data for 2007 showed 233

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738

applications, for a total of 18,867 pounds used on alfalfa, cotton, tomatoes, and potatoes.

soluble in aliphatic solvents; very soluble in aromatic solvents

Physical properties  Formula, C2H8NO2PS; MW, 128.1; MP, 45ºC; VP, 3.53  105 mm Hg at 25ºC; log P, 0.8; solubility in H2O, miscible in water; other solubilities: soluble in aliphatic chlorinated hydrocarbons, alcohols; slightly soluble in ether

Irritation Data  In the Draize assay, 78 and 87% liquid formulations of naled (Dibrom) caused corrosion (persistent edema, erythema, eschar, and necrosis 14 days after application). However, technical naled (97% concentration) caused only moderate (category III) irritation in the same irritation assay (see Table 28.6).

Irritation, sensitization, and illness data  The 40% EC is a minimal irritant in the Draize assay and is a nonsensitizer in the Buehler test. There was one case in the handler database – a case of dermatitis following accidental direct contact (see Table 28.6). O S OCH3 CH3O S N N P S CH2 CH3O Methidathion

Methidathion  Methidathion is also a potent cholinesterase inhibitor. California use data for 2007 showed 973 applications, for a total of 45,633 pounds used on forage, orchard, and row crops. Physical Properties  Formula, C6H11N2O4PS3; MW, 302.33; MP, 39°C; log P, 2.2; VP, 3.37  106 mm Hg; solubility in H2O, 187 mg/l; other solubilities: soluble in benzene, methanol, and xylene Irritation Data  A 22.6% EC caused moderate irritation in Draize assay. A 25% wettable powder caused minimal irritation. Sensitization Data  The 25% wettable powder was a nonsensitizer in the Buehler test. Illness Data  There were three cases associated with methidathion in the handler database, but there were no cases of dermatitis following accidental direct contact.

CH3O

P Naled

CH3O S P CH3O

O CH C Cl2 Br

Br

Naled  Naled is an OP that transforms to dichlorvos in the environment after application. Both compounds have relatively high vapor pressure and dissipate after application (Hall et al., 1997). California use data showed 2941 applications, with a total of 132,050 pounds used on grains, nurseries, cotton, vineyards, orchards, and safflowers. Physical Properties  Formula, C4H7Br2Cl2O4P; MW, 381; MP, 27°C; VP, 2.0  104 mm Hg at 20°C; log P, 1.38; solubility in H2O, 1.5 mg/l; other solubilities: slightly

O S CH2CH2 S C2H5 Oxydemeton-methyl

Oxydemeton-Methyl  Oxydemeton-methyl is a potent cholinesterase inhibitor. California agricultural use data showed 12,316 applications, for a total of 121,936 pounds used on row crops and nursery crops. Physical Properties  Formula, C6H15O4PS2; MW, 246.3; MP, 27°C; VP, 0.00195 mm Hg; water solubility, 1.5 mg/l; log P, 0.74; solubility in H2O, miscible; other solubilities: soluble in common organic solvents except petroleum ether Irritation, Sensitization, and Illness Data  There were no studies of dermal irritation available for review. A 50% formulation of oxydemeton-methyl was a nonsensitizer in the Buehler assay. There were two cases associated with oxydemeton in the handler database but none involving accidental direct contact (see Table 28.6). CH3CH2O CH3CH2O

O CH3O

Sensitization and Illness Data  No sensitization studies were available for review, but possible cases of contact sensitivity have been reported in the public domain literature (Edmundson and Davies, 1967). Cases of probable irritation (Mick et al., 1970) associated with naled have resembled the cases reported in pesticide handler database (e.g., cases 1988-2330 and 1988-942).

S P O

Ethyl parathion

NO2

CH3O CH3O

S P O

NO2

Methyl parathion

Parathion and Methyl Parathion  Parathion and methyl parathion are both potent cholinesterase inhibitors. Their skin effects have been a much less serious concern. Parathion has been off of the U.S. market since the early 1990s; methyl parathion is still in use. California use data for 2007 showed 1218 applications, for a total of 75,368 pounds used principally on walnuts; limited use was also reported for carrots, potatoes, orchards, and row crops. Physical Properties Parathion: formula, C10H14NO5PS; MW, 291.27; MP, 6.1°C; VP, 6.68  106; log P, 3.83; solubility in H2O,

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

11 mg/l; other solubilities: completely soluble in alcohols, esters, ethers, ketones, aromatic hydrocarbons, and animal and vegetable oils Methyl parathion: formula, C8H10NO5PS; MW, 263.21; MP, 35.5°C; VP, 3.50  106 mm Hg; log P, 2.86; solubility in H2O, 55 mg/l at 20°C; other solubilities: soluble in ethanol, chloroform, and aliphatic solvents Irritation Data  Technical formulations of liquid parathion (98% AI) and methyl parathion (80% AI) caused minimal irritation in the Draize assay. Sensitization Data  No dermal sensitization data were available for review for either product, but cases of possible sensitization were reported in the public domain literature for both. These included a possible case of sensitization in a vintner associated with parathion (with cross-reaction to methyl parathion, azinphos methyl, metathion, and oxydemeton methyl) reported by Pevny (1980). Svindland (1981) described an unusual topical reaction, resembling erysipeloid, in a gardener using a concentrated parathion/emsulfier spray who also had a cut finger. A case of erythema multiforme, associated with use of methyl parathion for structural pest control in India, has also been reported (Bhargava et al., 1977). Illness Data  There was one case associated with parathion in the pesticide handler database, occurring following an application accident (see Table 28.6). O CH3O CH3O

S P

S

CH2

Phosmet

N O

Phosmet  Phosmet is a moderately potent cholinesterase inhibitor. The nonphosphate “leaving group” portion of the compound somewhat resembles the phthalimido fungicides (captan and related compounds). California use data for 2007 showed 5230 applications, for a total of 421,109 pounds used for control of ectoparasites on large and small animals and for a vineyard, orchard, and nursery insecticide. Physical Properties  Formula, C11H12NO4PS2; MW, 317.3; MP, 72°C; VP, 4.9  107 mm Hg; log P, 2.78; solubility in H2O, 25 mg/l at 25°C; other solubilities (g/l at 25°C): acetone 650, benzene 600, kerosene 5, methanol 50, toluene and methyl isobutyl ketone 300, and xylene 250 Irritation and Sensitization Data  The 70% wettable powder caused minimal irritation in the Draize assay, but the 11.6% EC caused moderate irritation. The 5% dust was a nonsensitizer in the Buehler test.

739

Illness Data  There were five cases associated with phosmet in the handler database, including two cases of contact dermatitis of the hands following use of phosmet shampoo to control small animal ectoparasites (see Table 28.6). Cl

Cl O CH3O CH3O

Cl O CH CCl2

P

Tetrachlorvinphos

Tetrachlorvinphos  California use data for 2007 showed 667 pounds applied, principally for insect control in animal facilities. It is also used in pet flea collars. Total sales data showed 11,581 pounds sold during 2007. Physical Properties  Formula, C10H9Cl4O4P; MW, 317.3; MP, 97.5; VP, 4.20  108 mm Hg; log P, 3.53; solubility in H2O, 11 mg/l at 25°C; other solubilities (g/kg at 20°C): acetone 200, dichloromethane 400, and xylene 150 Irritation, Sensitization, and Illness Data  Several formulations of tetrachlorvinphos (75% wettable powder, fly control pellets with 1.2% AI, and 1.08% liquid) caused minimal irritation in the Draize assay. There were no dermal sensitization studies available for review and no cases associated with its use in the pesticide handler database.

28.2.2.2  Carbamates Cases of dermatitis associated with application carbamates are also reported in the public domain scientific literature (Bruynzeel, 1991; Vandekar, 1965). However, none of the currently registered carbamate insecticides appeared to be markedly irritant or consistently sensitize in animal test models. O H3C

HN

SCH3

C O N C C

Aldicarb

CH3

CH3

Aldicarb  California use data for 2007 showed 1405 applications, for a total of 115,031 pounds, principally on cotton. Minor uses included beans, pecan, nursery crops, and sugar beets. Physical Properties  Formula, C7H14N2O2S; MW, 190.27; MP, 99–100°C; VP, 9.75  105 mm Hg; log P, 1.13; solubility in H2O, 4.93 g/l at 20°C; other solubilities (g/kg at 25°C): acetone 350, dichloromethane 300, benzene 150, and xylene 150 Irritation, Sensitization, and Illness Data  For aldicarb, its systemic toxicity far outweighs its mild irritant effects.

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In the Draize assay, the 92% technical formulation caused serious systemic toxicity before the study could be completed. The 15% granular formulation caused no irritation. There were no dermal sensitization studies available for review. There was one case of dermatitis possibly associated with aldicarb in the pesticide handler data base (see Table 28.6), but there were no cases of contact dermatitis following accidental direct contact. O NH

O O

(7 and 6.3% granular and 0.5% liquid) caused minimal irritation (category IV) in the Draize assay. However, a 0.5% shampoo caused moderate irritation. The 90% dry flowable product is a nonsensitizer in the Buehler assay. However, among 30 Indian farmers with contact dermatitis evaluated by patch testing, several reacted to carbofuran and carbaryl. The relevance of the positive patch tests to the reported cases of dermatitis could not be determined from information reported in the study (Sharma and Kaur, 1990). Cases of irritant contact dermatitis in carbaryl applicators have occasionally been reported (six cases total in the applicator database; e.g., 1982-2634 and 1982-2703). O

O Bendiocarb

O

Bendiocarb  Bendiocarb was previously used extensively for structural pest control and had minimal use as a nursery insecticide in California during 2007. However, there were no active bendiocarb products listed on the U.S. EPA website. Physical Properties  Formula, C11H13NO4; MW, 223.23; MP, 129–130°C; VP, 3.45  105 mm Hg at 25°C; log P, 1.70; solubility in H2O, 260 mg/l at 25°C; other solubilities (g/kg at 25°C): chloroform 200, ethyl acetate 60–75, o-xylene 10, and p-xylene 11.7 Irritation, Sensitization, and Illness Data  There were no data on dermal irritation available for review, but bendiocarb was reported as a minimal irritant in the public domain literature. A mixture with 2.4% bendiocarb and 12% piperonyl butoxide was a nonsensitizer in the Buehler test. O O C NH CH3

CH2

O

CH2

Fenoxycarb  California use data for 2007 showed fenoxycarb had limited use, principally as a nursery insecticide. Physical Properties  Formula, C17H19NO4; MW, 301.3; MP, 53–54°C; VP, 6.5  109 mm Hg at 25°C; log P, 4.30; solubility in H2O, 6.0 mg/l at 20°C; other solubilities (g/l at 20°C): ethanol 510, acetone 770, toluene 630, n-hexane 5.3, n-octanol 130 Irritation, Sensitization, and Illness Data  A 40% wettable powder formulation of fenoxycarb caused severe irritation in the Draize assay. Other formulations, including a granular ant bait with 1% AI, an aerosol with 1.2% AI, a 23% liquid formulation, and technical fenoxycarb, caused only minimal (category IV) irritation in the Draize assay. No sensitization study was available for review, and no cases associated with fenoxycarb were identified in the pesticide handler database. O

Physical Properties  Formula, C12H11NO2; MW, 201.22; MP, 145°C; BP, 315°C; VP, 1.36E-06 mm Hg at 25°C; log P, 2.36; solubility in H2O, 110 mg/l at 25°C; other solubilities (g/kg at 25°C): dimethylformamide 400–450, dimethyl sulfoxide 400–450, acetone 200–300, cyclohexanone 200–250, isopropanol 100, and xylene 100 Irritation, Sensitization, and Illness Data  Technical ­ carbaryl (99%), as well as less concentrated products

C O CH2 CH3

Fenoxycarb

S

Carbaryl

Carbaryl  California use data for 2007 showed 2590 applications, for a total of 142,010 pounds used on grain, row, orchard, vineyard, and nursery crops; registered products in topical preparations for control of animal ectoparasites probably accounted for some additional use. 2007 data showed 323,069 pounds sold in California.

NH

NH

O Methiocarb

Methiocarb  California use data for 2007 showed 879 applications, for a total of 1737 pounds, used principally as a nursery insecticide, with limited use on avocados and citrus. Physical Properties  Formula, C11H15NO2S; MW, 225.3; MP, 119°C; VP, 2.7  107 mm Hg at 25°C; log P, 2.92; solubility in H2O, 27 mg/l at 20°C; other solubilities: dichloromethane 200 g/l, isopropanol 53 g/l, toluene 33 g/l, and hexane 1.3 g/l Irritation, Sensitization, and Illness Data  The 75% concentrate, 2% pellet, and 1% aerosol caused minimal irritation in the Draize assay.

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

O CH3

C

N

O

S

CH3

NH

C

CH3

Methomyl

Methomyl  California use data for 2007 showed 18,945 applications, for a total of 304,879 pounds applied to grains, forage, row crops, orchards, vineyards, cotton, and nursery plants. Physical Properties  Formula, C5H10N2O2S; MW, 162.21; MP, 78–79°C; VP, 2.7  107 mm Hg at 25°C; log P, 0.60; solubility in H2O, 58 g/l at 25°C; other solubilities: soluble in most organic solvents Irritation, Sensitization, and Illness Data  Technical methomyl (98.5%) and a high concentration (92.4%) caused no observable erythema or edema in the Draize assay. The technical material did not cause sensitization in the local lymph node assay. A fly bait product containing 0.025% muscalure and 1% methomyl was negative in the Buehler assay. There were two cases associated with methomyl in the pesticide handler database, both involving dermatitis following accidental direct contact (see Table 28.6, cases 1984-1512 and 1983-285). Oxamyl  California use data for 2007 showed 2080 applications, for a total of 44,843 pounds used on forage, grains, orchards, cotton, row crops, and nursery plants. Physical Properties  Formula, C7H13N3O3S; MW, 219.26; MP, 100–102°C; VP, 2.3  104 mm Hg at 25°C; log P, 0.47; solubility in H2O, 280 g/l at 25°C; other solubilities (g/100 ml at 25°C): acetone 67, ethanol 33, 2-propanol 11, methanol 144, and toluene 1 Irritation, Sensitization, and Illness Data  The 42% liquid caused minimal irritation in the Draize assay. A 10% liquid product caused sensitization in the local lymph node assay. There was one case associated with oxamyl in the pesticide handler database, but there were no cases of dermatitis following accidental direct contact. O O C

C

NH

CH3

CH3 O CH CH3 Propoxur

Propoxur  California use data for 2007 showed limited use, principally for structural pest control. California EPA data showed 6538 pounds sold in California for 2007.

741

Physical Properties  Formula, C11H15NO3; MW, 209.25; MP, 91.5°C; VP, 9.68  106 mm Hg at 20°C; log P, 1.52; solubility in H2O, 1.86 g/l at 30°C; other solubilities: soluble in methanol, acetone, and many organic solvents Irritation, Sensitization, and Illness Data  Technical propoxur (99.6% AI), a 70% wettable propoxur powder, a dog collar with a 10% concentration of propoxur, and two 0.5% ready-to-use formulations caused minimal irritation in the Draize assay. However, 14.6 and 19.6% ECs, as well as ready-to-use formulations containing 0.5 and 1% propoxur, caused moderate irritation. These somewhat contradictory findings suggest that the irritation caused by the less concentrated formulations was due to an inert ingredient rather than propoxur. A 1% ready-to-use formulation tested with the Buehler assay did not cause dermal sensitization. There were four cases associated with the use of propoxur in the pesticide applicator database (see Table 28.6). CH3

C

CH3

CH3

O C

O

Pyrethrin

CH2CH

CHCH

CH2

O

28.2.2.3  Pyrethrins The California Pesticide Label Database shows 583 currently registered pyrethrin formulations and an additional 3824 formulations previously registered.2 The cases associated with pyrethrins reflect their broad-scale use, almost always in mixtures with piperonyl butoxide and other synergists. (a) Pyrethrin I Physical Properties  Formula, C21H28O3; MW, 328.45; BP, 146–150°C, 0.005 mm Hg; VP, 2.03  105 mm Hg; log P, 5.9; solubility in H2O, 0.2 g/l at 25°C, but hydrolyzes in water, and the process is speeded by acid or alkali; other solubilities: soluble in alcohol, petroleum ether, kerosene, carbon tetrachloride, ethylene dichloride, and nitromethane Irritation Data  Dermal irritation studies with technical pyrethrins (57% concentration) show only transient erythema, disappearing by 72 h (category IV in the Draize assay). Irritation suffered by users (e.g., case 1990-2621, Table 28.6) may be due to pyrethrins but could also be caused by petroleum distillates (common inert ingredients in ready-to-use formulations) or synergists contained in ­ formulated products. For 2

Search date 2/9/2009: Available at http://www.cdpr.ca.gov/docs/label/ labelque.htm.

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742

example, a concentrated formulation of piperonyl butoxide (92%) caused severe irritation in the Draize test. Sensitization Data  Five products (0.124–9% pyrethrins, mixed with other insecticides and synergists) caused sensitization in the Buehler test. An additional product (0.4% pyrethrins, 0.025% deltamethrin) was positive in the GPMT. Twelve products (0.05–5% pyrethrins, including mixtures with synergists and other insecticides) caused no sensitization in the Buehler test. It is not clear that pyrethrins are clinically significant sensitizers. The active sensitizers in older formulations were probably sesquirtepene lactone contaminants of crude pyrethrum extracts (Feinberg, 1934). However, most pyrethrin products are labeled as potential sensitizers. Illness Data  Because most pyrethrin formulations contained synergists (piperonyl butoxide and N-octyl-bicycloheptene dicarboximide, described later), the 40 pyrethrin cases involving applicators were coded as mixtures in the California illness registry. The cases consequently did not match the criterion for inclusion in the database given previously, but they otherwise resembled the handler cases.

28.2.2.4  Pyrethrin Synergists O

CH2CH2CH3

O

CH2OCH2CH2OCH2CH2OC4H9 Piperonyl butoxide

Piperonyl Butoxide  Piperonyl butoxide (PBO) is a synergist used with formulations of pyrethrins and pyrethroids, but it is not chemically related to either group. Physical Properties  Formula, C19H30O5; MW, 338.43; BP, 180°C, 1 mm Hg; VP, 2.6  107 mm Hg; log P, 4.75; solubility in H2O, 14.3 g/l at 25°C; other solubilities: miscible with methanol, ethanol, benzene, freons, petroleum oils, and other organic solvents Irritation, Sensitization, and Illness Data  In the Draize assay, a 91% technical formulation was classified as a moderate (category III) irritant and a 92% technical formulation was a severe (category II) irritant. Several other technical formulations (92–100% PBO) only showed minimal irritation. Ready-to-use products with 25 and 0.49% PBO were nonsensitizers in the Buehler assay. Cases reported in the handler database also had simultaneous exposures to pyrethrins and inert ingredients. Most involved ready-to-use formulations containing PBO in concentrations less than 1% and not expected to cause irritation in handlers.

O N O n-octyl-bicycloheptene-dicarboximide

N-octyl-bicycloheptene dicarboximide (NOBD) NOBD is also used as a pyrethrin synergist in hundreds of formulations. Physical Properties  Formula, C17H25NO2; MW, 275.4; BP, 158.2°C, 2 mm Hg; VP, 1.8  105 mm Hg; log P, 3.7; solubility in H2O, 13.7 g/l at 25°C; other solubilities: miscible with most organic solvents including petroleum oils and fluorinated hydrocarbons Irritation and Sensitization Data  No study evaluating the isolated effect of NOBD was available for review (see Table 28.6), but there were several studies involving mixtures with active pesticidal compounds. There was no study available on the capacity of NOBD to cause sensitization. A formulation containing 20% NOBD, 70% diethyl toluamide, 5% isochromyl cinchonerate, and 5% bis butenylene tetrahydro furfural and a mixture containing 13.4% NOBD, 1.0% prallethrin, and 13.34% cyphenothrin produced only transient (category IV) irritation in the Draize assay. Nevertheless, a product containing 12.5% NOBD and 3.02% prallethrin produced moderate (category III) irritation in the Draize assay. By inference, the more intense irritation reported in this study was attributable to the 3% concentration of prallethrin rather than to NOBD. Illness Data  Cases involving NOBD reported to the hand­ ler database all involved mixtures with pyrethrins and other compounds.

28.2.2.5  Synthetic Pyrethroids Synthetic pyrethroids have replaced organophosphates for some structural pest control and also agricultural applications. The effects of synthetic pyrethroids on the sodium channels of cutaneous nerve endings may cause paresthesias at levels of exposure that do not provoke visible erythema (Lisi, 1992). The standard Draize imitation study may be a poor means for evaluating such purely symptomatic endpoints. An alternative animal test developed by Cagen evaluates the sensory effect of pyrethroids through observations of grooming behavior focused on the site(s) of applied test material. The behavioral test demonstrated direct effects on grooming behavior for 4 h after pyrethroid

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

application and increased response to other chemical irritants (oil of mustard in the test model) for 24 h after application (Cagen et al., 1984). Illness Data  There were 40 cases associated with synthetic pyrethroids in the pesticide handler database, including 15 (38%) associated with type I pyrethroids and 25 (62%) with type II compounds. There were an additional 66 cases that involved mixtures with synergists (similar to the formulations of pyrethrins described previously). These cases did not meet the criteria for inclusion in the handler database. (a) Type I Pyrethroids These contain two cis/trans isomeric sites and may have as many as four isomers with the ability to stimulate cutaneous nerves in the human epidermis (Flannigan and Tucker, 1985a; Flannigan et al., 1985a,b; Gammon, 1985; Gammon and Casida, 1983; Tucker et al., 1984). These sensory effects are not evaluated in the Draize assay for cutaneous irritation. CH3 C CH3

C CH3

CH3

O

CH

O

CH3

CH2CH

CH2

O

d-allethrin

Allethrin  California EPA data indicated 4781 pounds of allethrin were sold in California in 2007. Physical Properties  Formula, C19H26O3; MW, 302.4; BP, 281.5°C, 760 mm Hg; VP, 1.2  106 mm Hg; log P, 4.78; solubility in H2O, 4.6 mg/l at 25°C Irritation Data  In the Draize assay, one formulation of technical d-allethrin, a cis/trans mixture, caused corrosion (category I irritation). Two different tests on a 96.1% technical formulation showed minimal (category IV) irritation in the Draize assay, and tests on a 92.1% liquid formulation and a similar 92.4% liquid also showed minimal (category IV) irritation. Draize tests on seven ready-to-use end products (including mixtures with resmethrin, phenothrin, dipropyl isocinhomeronate, PBO, and NOBD) all showed minimal irritation. Tests on five end-use products – one containing only allethrin, two with mixtures containing permethrin, one with a mixture of PBO and NOBD, and one with a mixture containing allethrin, NOBD, and chlorpyrifos – showed moderate irritation. Sensitization and Illness Data  A Buehler sensitization study was negative on a dilute end-use product (a mixture of allethrin, cypermethrin, piperonyl butoxide, and petroleum distillates).

743

Case 1994-438 involved irritation on direct accidental exposure to a formulation of allethrin and piperonyl butoxide (see Table 28.6; not included in the pesticide handler database). O CF3

C O

C CH

Cl Bifenthrin

Bifenthrin Physical Properties  Formula, C23H22ClF3O2; MW, 422.87; MP, 69°C; VP, 1.8  107 mm Hg; log P, 6.00; solubility in H2O, 0.1 mg/l; other solubilities: soluble in methylene chloride, chloroform, acetone, ether, and toluene; slightly soluble in heptane and methanol Irritation Data  Thirteen products (0.18–94% AI) caused minimal irritation in the Draize assay. Three products (2.4– 13.2% AI) caused moderate irritation and 3 caused severe irritation (13–24.9% AI). Sensitization Data  Fourteen products (0.184–88.3% AI) were negative in the Buehler test. Two products (4% EC and 26% liquid) were both positive in the same assay. CCl2

CH

O

O C O CH2

CH3

CH3 Permethrin

Permethrin  California use data for 2007 showed 39,343 applications, for a total of 413,837 pounds used on forage, orchards, row crops, and nursery plants. Additional use occurs with the application of home-use products (typically mixtures with pyrethrins, synergists, and other pyrethroids). A 1% lotion is used as a pharmaceutical treatment for pediculosis capitis and a 5% cream as a treatment for scabies (EPOCRATES, 2009). Physical Properties  Formula, C21H20Cl2O3; MW, 391.29; MP, 34–35°C; BP 290°C, 760 mm Hg; VP, 2.18  108 mm Hg; log P, 6.50; solubility in H2O, 6.00  103 mg/l at 20°C; other solubilities: soluble in most organic solvents except ethylene glycol Irritation Data  Five products (0.5 and 0.72% sprays, two 10% liquids, and a 36.8% EC) caused severe irritation in the Draize assay. Four additional products (4.6% liquid, 2.5% permethrin liquid, 0.18% permethrin spray mixture, a 0.5% permethrin repellant) caused moderate irritation. Forty-five products (0.2% spray to 95.2% solid technical) caused minimal irritation.

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Sensitization Data  Three products (95.6% liquid technical material, 92% liquid technical, and 25% wettable powder) caused sensitization in the GPMT. Four products (30.2% liquid, 10% emulsion, 4.6% liquid, and 4% liquid and 0.2% spray mixture) caused sensitization in the Buehler test. A 45% liquid caused sensitization in the LLNA. Fifteen products (0.2–65% liquid) were negative in the Buehler test. Illness Data  There were eight cases associated with permethrin in the handler database. These included case 19921381, irritation of the hands and forearms that gradually occurred over the course of a day while mixing permethrin as material accumulated underneath his gloves and shirt. CH3

C

O

CH

CH3

C O

O

Phenothrin

Phenothrin  California data for 2007 showed limited use of phenothrin. California EPA data showed 95,233 pounds sold in California during 2007. Physical Properties  Formula, C23H26O3; MW, 350.46; BP, 290°C; VP, 1.43  107 mm Hg; log P, 7.540; solubility in H2O, 0.0097 mg/l at 25°C; other solubilities (g/ml at 25°C): methanol 5.0 and hexane 4.96 Irritation Data  In the Draize assay, a 10% formulation of phenothrin and six end-use mixtures (with NOBD, isopropanol, quaternary ammonium compounds, and tetramethrin) caused minimal (category IV) irritation. Three formulations, with 0.1–0.4% phenothrin mixed with NOBD, quaternary ammonium compounds, or tetramethrin, caused moderate irritation in the Draize assay. Sensitization Data  Buehler assays on an end-use formulation (with 0.4% phenothrin and 1.5% NOBD) and a 10% concentrate were both negative for sensitization. A readyto-use product containing 0.1% phenothrin and 0.05% tetramethrin showed evidence of sensitization (based on slightly increased erythema between challenge and rechallenge exposures). Illness Data  There were two cases associated with phenothrin in the handler database, including a typical case of contact dermatitis following direct accidental exposure (see Table 28.6). CH3 CH3

C

CH

O C

CH3

CH2 O

CH2

CH3 Resmethrin

O

Resmethrin  California data showed limited use of resmethrin for 2007, principally for nursery plants. Physical Properties  Formula, C22H26O3; MW, 338.4; MP, 56.5°C; BP, decomposes at 180°C; VP, 1.13  108 mm Hg; log P, 5.43; solubility in H2O, 0.0379 mg/l at 25°C; other solubilities: very solvent in xylene and aromatic petroleum hydrocarbons; solubility in kerosene 10% Irritation Data  Technical resmethrin (88% AI) caused moderate irritation in the Draize assay. Formulations containing 0.25, 0.58, 0.716, 1, and 3% resmethrin and two formulations (0.08–0.2% resmethrin, mixed with allethrin, pyrethrins, and PBO) caused minimal irritation. In an outlying result, a 3.41% formulation of resmethrin in petroleum distillate caused corrosive (category I) irritation in the Draize assay. Sensitization Data  A product containing 3% resmethrin aqueous concentrate caused sensitization in the Buehler assay. Six products (0.05–4% AI) caused no sensitization. Illness Data  There were three cases included in the handler database, including one case of topical irritation after accidental direct contact (case 2000-122, Table 28.6). CH3 CH3

C CH

C O CH3

O

O

CH3 Tetramethrin

CH2 O

Tetramethrin  California data showed limited agricultural use of tetramethrin for 2007. It is still a common ingredient in ready-to-use household insect sprays. California EPA data showed 13,126 pounds sold in California for 2007. Physical Properties  Formula, C19H25NO4; MW, 331.41; MP, 68–70°C; BP, 180–190°C, 0.1 mm Hg; VP, 7.1  106 mm Hg; log P, 4.73; solubility in H2O, 1.83 mg/l at 25°C; other solubilities: methanol (53 g/kg), hexane (20 g/kg), xylene (1 g/kg), acetone, and toluene Irritation, Sensitization, and Illness Data  A 21% formulation of tetramethrin caused minimal (category IV) irritation in the Draize assay. There were no other studies of the isolated irritant effects of tetramethrin. Three formulations containing 0.2–12% tetramethrin mixed with permethrin, esfenvalerate, or resmethrin caused minimal (category IV) irritation in the Draize assay. Formulations containing 0.2–16.7% tetramethrin mixed with resmethrin, permethrin, PBO, or esfenvalerate caused moderate (category III) irritation.

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

A formulation with 0.15% tetramethrin, 0.375% permethrin, and 0.75% PBO was negative in the Buehler sensitization assay. There were no cases with isolated exposure to tetramethrin in the handler database. (b) Type II Pyrethroids These contain as many as three isomeric sites, and most contain a cyano group attached near the ester linkage. They are relatively more potent systemic toxins than the type I pyrethroids and cause a greater degree of paresthesia in experimental studies on human volunteers. O

Cl C

CH

CH

Cl

CH C

CN O

CH

O

C CH3

F

CH3 Cyfluthrin

Cyfluthrin  California agricultural use data for 2007 showed 8646 applications, for a total of 25,707 pounds, used on forage, orchards, row crops, cotton, vineyards, and corn and for structural pest control. Physical Properties  Formula, C19H25NO4; MW, 331.41; MP, 56.5°C; BP, 180–190°C, 0.1 mm Hg; VP, 7.1  106 mm Hg; log P, 4.73; solubility in H2O, 1.83 mg/l at 25°C; other solubilities: methanol (53 g/kg), hexane (20 g/kg), xylene (1 g/kg), acetone, and toluene Irritation Data  A 20% wettable powder caused minimal (category IV) irritation in the Draize assay. Similar results were found with six dilute end-use formulations containing 0.1% cyfluthrin (including four formulations containing mixtures of prallethrin, pyrethrins, PBO, and NOBD). Formulations associated with moderate (category III) irritation in the Draize assay included a 1% pour-on treatment for cattle ectoparasites, a 6% aqueous concentrate, and two different ECs containing 25% cyfluthrin. Severe irritation (category II) associated with a dilute end-use product containing 0.1% cyfluthrin, tetramethrin, and PBO was an outlying result. Sensitization Data  Three products were evaluated for sensitization in the Buehler assay. A mild sensitization response was observed for a recently registered 25% EC. No sensitization was observed in an earlier study of a similar 25% EC formulation submitted by a different registrant. Assays performed on a 0.1% granular formulation and a ready-to-use aerosol formulation containing 0.1% cyfluthrin, pyrethrins, and PBO were both negative for sensitization. Illness Data  There were 12 cases associated with cyfluthrin in the pesticide handler database (see Table 28.6). In a case of contact dermatitis following direct contact, the most

745

prominent reported symptom was erythema, rather than paresthesia or purely sensory irritation (case 1999-497). CF3

O C

CH

CH

Cl

CH

C

CN O

CH

O

C CH3

CH3

Lambda-cyhalothrin

Cyhalothrin  Cyfluthrin and cyhalothrin are similar in structure, differing principally in substitution of a trifluoro­ methane in cyhalothrin for a chlorine atom in cyfluthrin and substitution of a fluorine on one of the phenoxybenzyl ether aromatic rings. The cyhalothrin use profile is similar to that for cyfluthrin. California data for 2007 showed 27,088 applications, for a total of 31,633 pounds applied, used on forage, orchards, row crops, cotton, vineyards, and corn and for structural pest control. Physical Properties  Formula, C23H19ClF3NO3; MW, 449.86; MP, 49.2°C; BP, 187–190°C, 0.2 mm Hg; VP, 7.5  109 mm Hg; log P, 6.8; solubility in H2O, 5.0  103 mg/l at 25°C; other solubilities: 500 g/l at 20°C in acetone, dichloromethane, methanol, diethyl ether, ethyl acetate, hexane, and toluene Irritation Data  A 25% microencapsulated formulation of cyhalothrin caused minimal irritation in the Draize assay. Minimal irritation was also observed for other formulations (12.7% EC, 11.4% microencapsulated formulation, 9.7% aqueous concentrate, 0.05% ready-to-use aerosol formulation, 0.04% granules, and a 0.03% ready-to-use aerosol). Moderate (category III) irritation was associated with several cyhalothrin products, including a 9.53% microencapsulated formulation and two 10% wettable powders. Sensitization Data  The Buehler assay showed weak or equivocal sensitization in three formulations (10% cyfluthrin and 13% PBO in a cattle ear tag and a 24% granular formulation). A 9.7% aqueous concentrate was a weak sensitizer in the GPMA. Two end-use formulations (0.03 and 0.05%) were negative in the Buehler assay. Illness Data  There were two cases associated with handling cyhalothrin, including a case of dermatitis and conjunctivitis following direct contact (case 2003-497, Table 28.6). Cl C

CH

CH

Cl

CH

O

CN

C

O CH

C CH3

CH3 Cypermethrin

O

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746

Cypermethrin  California agricultural use data for 2007 showed 1761 applications, for a total of 336,826 pounds used on forage, orchards, nursery plants, row crops, cotton, vineyards, and corn and for structural pest control. Physical Properties  Formula, C22H19Cl2NO3; MW, 416.30; MP, 80.5°C; VP, 3.07  109 mm Hg; log P, 6.60; solubility in H2O, 4  103 mg/l at 25°C; other solubilities: acetone, chloroform, cyclohexanone, xylene 450, ethanol 337, hexane 103 (all in g/l at 20°C) Irritation Data  A 26% EC caused severe (category II) irritation in the Draize assay. Six formulations (including two 25.3% ECs, an 18.1% EC, and three formulations containing 1% cypermethrin) caused moderate irritation in the Draize assay. Nine formulations (including a 24.8% EC, 17.1% concentrate in oil, a cattle ear tag with 16% cypermethrin and 20% PBO, two foggers with 1.7% cypermethrin, two aerosols with 1% or less AI, and two mixtures with pyrethrins and synergists containing 1% cypermethrin) caused minimal irritation in the Draize assay. Sensitization Data  A Buehler study on a 1% spray for equine ectoparasites was negative for sensitization. However, a formulation with 0.824% cypermethrin in triethylene glycol showed evidence of mild sensitization in the same assay. Public domain literature showed that cypermethrin was negative in both the standard in vivo local lymph node assay and an in vitro variation. The same publication showed that cypermethrin was a sensitizer in the GPMT (a test that eliminates the effect of dermal absorption). Illness Data  There were five cypermethrin cases in the handler database, including dermatitis associated with accidental transfer of cypermethrin from the hands to the genitalia (case 1988-2388, Table 28.6). O

Br C

CH

CH

Br

CH

C

CN O

CH

O

C CH3

CH3 Deltamethrin

Deltamethrin  California data for 2007 showed 1419 applications, for a total of 20,581 pounds applied used on forage, orchards, row crops, cotton, nursery plants, and corn and for structural pest control.

Irritation, Sensitization, and Illness Data  The technical formulation of deltamethrin (98% AI), as well as 10 less concentrated formulations (0.01–4.95% deltamethrin), caused minimal (category IV) irritation in the Draize assay. Two formulations (an 11.7% gel and a 2.86% EC) caused severe (category III) irritation in the animal assay. Two dilute formulations (0.01% deltamethrin) were nonsensitizers in the Buehler assay. There was one case associated with deltamethrin in the handler database (2001-799) – a case of contact dermatitis following accidental direct exposure. CH3 Cl

CH3 CH CH

CN C

O

O

CH

O Esfenvalerate

Esfenvalerate  Esfenvalerate is a stereoisomer of fenvalerate. California use data for 2007 showed 21,052 applications, for a total 42,780 pounds used on orchards, row crops, and nursery crops. Physical Properties  Formula, C25H22ClNO3; MW, 419.9; BP, 151–167°C; MP, 59–60.2°C; VP, 1.5  109 mm Hg; log P, 6.22; solubility in H2O, 2.0  103 mg/l at 25°C; other solubilities (g/kg at 25°C): xylene, acetone, chloroform, ethyl acetate, dimethylformamide, dimethyl sulfoxide 600, hexane 10–50, and methanol 70–100 Irritation, Sensitization, and Illness Data  A 24.6% EC, a 9.53% suspension, a 3.48% concentrate, and two readyto-use formulations with less than 1% AI caused moderate irritation and a single 0.443% emulsifiable formulation caused severe irritation in the Draize assay. Seven formulations, including a 35% wettable power, an 18% liquid formulation, and five formulations with less than 1% AI, caused minimal irritation. Six formulations tested in the Buehler assay (18% liquid, 9.53% suspension, and four formulations with 1% AI) were negative for sensitization. There were two cases associated with esfenvalerate in the handler data, including a typical case of irritation following direct contact (case 2000-534, Table 28.6). CH3 Cl

CH3 CH CH

C

CN O

O

CH

O

Physical Properties  Formula, C22H19Br2NO3; MW, 505.21; MP, 101–102°C; VP, 1.5  10-8 mm Hg; log P, 6.20; solubility in H2O, 0.002 mg/l; other solubilities: soluble in ethanol, acetone, and dioxane

Fenvalerate

Fenvalerate  California agricultural use data showed use of fenvalerate in 2007.

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

Physical Properties  Formula, C25H22ClNO3; MW, 419.90; BP, decomposes; VP 1.5  109 mm Hg; log P, 6.20; solubility in H2O, 2.4  102 mg/l at 22°C; other solubilities (g/l at 20°C): acetone 450, chloroform 450, methanol 450, and hexane 77 Irritation, Sensitization, and Illness Data  A 10.5% water dispersable formulation and a mixture with 0.4% fenvalerate, pyrethrins, PBO, and NOBD caused minimal irritation in the Draize assay. A mixed formulation with 0.5% fenvalerate, PBO, and NOBD caused severe irritation, and a mixture of 0.4% fenvalerate, chlorpyrifos, and DDVP caused moderate irritation in the animal assay. No sensitization assays were available for review, although several were available for the related compound, esfenvalerate. There was one case associated with fenvalerate in the handler database—a case of contact dermatitis following direct accidental exposure (see Table 28.6).

28.2.2.6  Organochlorines

747

more rapid environmental and metabolic degradation. California use data for 2007 showed 1124 applications, for a total of 75,969 pounds used on orchards, row crops, cotton, vineyards, and nursery plants. Physical Properties  Formula, C14H9Cl5O; MW, 370.47; MP, 77–78°C; BP, 180°C, 1 mm Hg; VP, 3.98  107 mm Hg; log P, 4.28; solubility in H2O, 1.2 mg/l at 24°C; other solubilities (g/l at 20°C): soluble in most aliphatic and aromatic solvents Irritation, Sensitization, and Illness Data  A 50% soluble powder and a 42% dicofol EC caused minimal (category IV) skin irritation in the Draize assay. A separate study of the 42% EC showed moderate irritation. A Buehler study of the same formulation showed no evidence of sensitization. The handler database contains four cases associated with dicofol, including a report of irritant contact dermatitis following prolonged contact with dicofol occluded against the skin (84-954 and 84-1454). Cl

C Cl

Cl

DDT

Cl Cl

Cl

Dienochlor

Dichloro-Biphenyl-trichloroethane Dichloro-biphenyl-trichloroethane (DDT) is the model organochlorine compound but is not used legally anywhere in the world, except in public health vector control campaigns. It has a remarkably high octanol/water partition coefficient and high fat solubility. Organochlorines still registered in the United States are those with the shortest environmental and biological half-lives. Physical Properties  Formula, C14H9Cl5; MW, 354.49; BP, 260°C; MP, 108.5°C; VP, 1.6  107 mm Hg; log P, 6.91; solubility in H2O, 5.5  103 mg/l at 25°C; other solubilities: high solubility in fat (100,000 ppm); g/100 ml solvent: acetone 58, benzene 78, benzyl benzoate 42, carbon tetrachloride 45, chlorobenzene 74, cyclohexanone 116; 2 g/100 ml 95% alcohol: ethyl ether 28, gasoline 10, isopropanol 3, kerosene 8–10, morpholine 75, peanut oil 11, pine oil 10–16, tetralin 61, tributyl phosphate 50 C

Cl

Cl

C H

Cl

Cl Cl

Cl3

Cl3

C

Cl

Dienochlor  There are no currently registered formulations of dienochlor. The 26 previously registered formulations were used for mite control or control of mealy bugs. Physical Properties  Formula, C10Cl10; MW, 474.64; MP, 122–123°C; BP, 250°C (decomposes); VP, 2.18   106 mm Hg; log P, 3.23; solubility in H2O, 2.50   102 mg/l at 20°C; other solubilities (g/l at 20°C): slightly soluble in ethanol, acetone, and aliphatic hydrocarbons; moderately soluble in aromatic hydrocarbons Irritation, Sensitization, and Illness Data  There were no irritation or sensitization data available for dienochlor. There were three cases in the handler database but no cases of direct contact dermatitis following accidental contact. A patch study published in the public domain literature demonstrated that skin metabolism of dienochlor is associated with brown discoloration of the skin, similar to that produced by a standard patch test material, balsam of Peru (O’Malley et al., 1995; Penagos et al., 2000). A case of sensitization related to benomyl and dienochlor was reported in The Netherlands (van Joost et al., 1983).

OH

Cl

Dicofol

Dicofol  Dicofol is an agricultural insecticide with three currently registered products and 127 formulations prev­ iously registered in California. It is structurally similar to DDT but contains a central hydroxyl group that allows

Cl

Cl

Cl

O S O

Cl Cl

Endosulfan

O

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748

Endosulfan  California use data for 2007 showed 1770 applications, for a total of 52,403 pounds used on forage, row crops, vineyards, and nursery plants and for landscape maintenance. Physical Properties  Formula, C9H6Cl6O3S; MW, 406.93; MP, 106°C; BP, 106°C at 0.7 mm Hg (with partial decomposition); VP, 6.2  106 mm Hg; log P, 3.83; solubility in H2O, 0.53 mg/l at 20°C; other solubilities: soluble in xylene, kerosene, chloroform, acetone, and alcohol Irritation, Sensitization, and Illness Data  A 50% wettable powder and a 33.7% EC caused minimal irritation in the Draize assay. The 33.7% EC formulation caused sensitization in the GPMT. A case of erythema multiforme associated with the use of pyrethrins and endosulfan was also reported from Spain. Patch testing showed positive reactions to both materials ( to 2% pyrethrums in petrolatum,  to endosulfan 1% aqueous solution); a skin biopsy showed a lymphocytic dermal infiltrate. The authors believed that the  reaction was not significant because of the absence of prior reports of sensitization associated with endosulfan (Garcia-Bravo et al., 1995). There were no cases associated with endosulfan in the pesticide handler database. Cl Cl

Cl

Cl

Cl Cl Lindane

Lindane  Lindane was used historically for control of ectoparasites on cattle and for control of insects on a variety of commercial crops. California data for 2007 showed limited agricultural use. Its pharmaceutical use has also been prohibited in California, but it is still registered elsewhere in the United States (O’Malley, 2007). Physical Properties  Formula, C6H6Cl6; MW, 290.83; MP, 112.5°C; BP, 323.4°C at 760 mm Hg (with partial decomposition); VP, 4.20  105 mm Hg; log P, 3.72; solubility in H2O, 7.3 mg/l at 25°C; other solubilities (g/l at 20°C): acetone 200, methanol 29–40, xylene 250, ethyl acetate 200, n-heptane 10–14 Irritation, Sensitization, and Illness Data  There were no dermal irritation studies available for review, but a 20% formulation used for control of borers and leaf miners is labeled as a skin irritant. It is also a sensitizer in the Buehler assay. Dermatitis has also been reported among workers in lindane manufacturing operations, but the reported cases

were possibly attributable to precursors and by-products not typically found in commercial formulations of lindane (A. Smith, 1991). As discussed previously, topical permethrin products (EPOCRATES, 2009) are the recommended first-line treatment for both pediculosis and scabies. Although the agricultural products may contain as much as 40% lindane, post-treatment dermatitis has also occasionally occurred in patients treated for scabies with 1% formulations of lindane. The extensive series reported by Farkas (1983) also contained cases reacting to a 20% scabicidal formulation of sulfur. The handler database contained three cases of contact dermatitis following direct accidental exposure to lindane, all consistent with irritant reaction (see case example 1986309, Table 28.6). CCl3 CH3O

C

OCH3

H Methoxychlor

Methoxychlor  Methoxychlor is a DDT analog, differing principally in the two opposing methoxy substituents in place of two chlorine atoms present in DDT. There are no products containing methoxychlor currently registered in the United States, but 260 formulations were previously registered. Its range of applications was similar to that for DDT, but it had a markedly shorter environmental and biological half-life (A. Smith, 1991). Physical Properties  Formula, C16H15Cl3O2; MW, 345.65; MP, 87°C; BP, 346°C; VP, 2.58  106 mm Hg; log P, 5.08; solubility in H2O, 0.1 mg/l at 25°C; other solubilities: moderately soluble in alcohol and petroleum oils; readily soluble in most aromatic solvents Irritation, Sensitization, and Illness Data  No animal data are on file for either sensitization or irritation studies, but methoxychlor is identified as slightly irritant in public domain literature (National Library of Medicine, 2009). There were no cases associated with methoxychlor in the handler database.

28.2.2.7  Biological Insecticides and Repellants The biological products discussed next exceed the molecular weight limits for expected skin irritants in the SICRET model. For example, the model predicts that protein products (group CNS) with molecular weight greater than 620 will not cause irritation. For large multicyclic compounds containing only carbon, hydrogen, and oxygen (group C), the corresponding upper limit of molecular weight is 350. Bacillus Thuringiensis  The California pesticide label database showed 369 Bacillus thuringiensis (Bt) products,

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

including 43 active registrations for 17 strains of Bt and 5 varieties of Bt endotoxin. A complete description of Bt use patterns is beyond the scope of this chapter. California use data for a representative AI, Bacillus thuringiensis (Berliner), subsp. aizawai, gc-91 protein, showed 2529 applications, for a total of 20,474 pounds used on forage, orchards, row crops, root vegetables, berries, and nursery plants. Physical Properties  Wild-type Bt kuristaki contains a plasmid encoding a delta-endotoxin (sometimes also described as Bt exotoxin) with a molecular weight of 130 kDa. The 130-kDa protein is a pro-toxin, broken down by insect intestinal proteases into an active compound with a molecular weight of approximately 60 kDa (Lightwood et al., 2000). Based on its size, the SICRET model predicts the endotoxin to be nonirritating to the skin.

in H2O, 260 mg/l at 25°C; other solubilities: readily soluble in ethanol, diethyl ether, acetone, and chloroform; insoluble in hexane Irritation, Sensitization, and Illness Data  The 40% technical material and seven other formulations (containing 0.15–4.5% AI ) caused minimal (category IV) irritation and two 3% formulations caused moderation (category III) irritation in the Draize assay. Buehler sensitization studies of four formulations (0.15–1.9% AI) were negative. A guinea pig maximization study of a 1.8% EC was positive for sensitization. There were no cases associated with azadiracthin in the handler database.

HO

Irritation, Sensitization, and Illness Data  For Bt strain Berliner, five formulations (0.5–8.0% Bt, including two formulations containing 25–50% elemental sulfur) caused minimal (category IV) irritation. Two formulations (3.2 and 6.4% Bt) caused moderate (category III) irritation in the Draize assay. A formulation containing 0.436% Bt, tested as a dried concentrate containing 52% AI, was negative for sensitization in the Buehler assay. There was one case associated with Bt in the handler database but no cases of contact dermatitis following direct accidental contact (see Table 28.6).

OCH3 OCH3

CH3

O

CH3

O O CH3

O

O

O

Avermectin 1b O OH O

CH3 H CH3

O

CH3 OCH3

H

CH3

OH

O

O H

749

CH3

O

O

O

H

OH

HO

Abamectin  Abamectin is a mixture of avermectin Bla and Blb used as an ant and cockroach bait and sometimes for nursery and agricultural pest control.

CH3

O

H

CH3 O

O

CH3

H O H3C

O

O O

O CH3

Azadirachtin  Azadirachtin is a triterpenoid derived from neem oil. California use data for 2007 showed 11,556 applications, for a total of 2220 pounds used on forage, orchards, row crops, vineyards, mushrooms, nursery plants, root vegetables, and berries. Physical Properties  Formula, C35H44O16; MW, 720.7; MP, 154–158°C; VP, 2.7  1011 mm Hg; log P, 1.09; solubility

Physical properties  Formula, C48H72O14.C47H70O14; MW, 873.09; MP, 150–155°C; VP, 1.5  109 mm Hg; log P, NA; solubility in H2O, 10 g/l at 21°C; other solubilities (g/L at 21°C): acetone 100, n-butanol 10, chloroform 25, cyclohexane 6, ethanol 20, isopropanol 70, kerosene 0.5, methanol 19.5, and toluene 350 Irritation, Sensitization, and Illness Data  Nine formulations of abamectin (0.01–46.3% AI) caused minimal irritation in the Draize assay. Five formulations (0.01–2.21% AI) caused moderate irritation. As with azadirachtin, sensitization studies with the Buehler method were negative (for six products containing 0.011–49.6% AI). However, abamectin was positive for sensitization in the GPMT (with a 1.8% EC formulation). There were seven cases associated with abamectin, including three cases of contact dermatitis following direct accidental exposure (see Table 28.6).

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28.2.2.8  Miscellaneous Insecticides and Repellants CH3

CH3 CH3

N

CHN

Amitraz

CH

N

CH3

CH3

Amitraz  The aniline derivative amitraz is used for ectoparasites in veterinary practice and for pear psylla control on pears and for whitefly on cotton. Current use is limited because of concerns about potential reproductive effects. A total of 66,439 pounds were reported used in California during 1997 compared with 12 pounds in 2006 and none in 2007. Physical properties  Formula, C19H23N3; MW, 293.45; MP, 86–87°C; VP, 2.0  106 mm Hg; log P, 5.50; solubility in H2O, 1 mg/l at 25°C; other solubilities: soluble in most organic solvents Irritation, Sensitization, and Illness Data  Five EC formulations of amitraz (12.5–20% AI) caused moderate irritation in the Draize assay. A flea collar with 9% amitraz produced only mild irritation in the assay. An amitraz product formulated as a 5% wettable powder proved to be a sensitizer in the GPMT. There were no cases associated with amitraz in the pesticide handler database. Borates (Boric Acid and Sodium Borate)  California EPA data showed 635,317 pounds of borates sold in California in 2007. Use reporting for 2007 showed that most borate use was for structural pest control. Physical Properties Sodium borate: formula, B4Na2O7; MW, 201.22; MP, 743°C, decomposes; VP, approximately 0 mm Hg; solubility in H2O, 31 g/l at 25°C; other solubilities: 0.60 g/100 g acetone, insoluble in alcohol Boric acid: formula, BH3O3; MW, 61.833; MP, 170.9°C; VP, negligible at 20°C; log P, 0.175; solubility in H2O, 47.2 g/l at 20°C; other solubilities: lycerol 17.5% at 25°C; ethylene glycol 18.5% at 25°C; in methanol 173.9 g/l at 25°C; in ethanol 94.4 g/l at 25°C; in acetone 0.6% at 25°C; and ethyl acetate 1.5% at 25°C Irritation, Sensitization, and Illness Data  Borates (including borax and hydrated octaborates) are nonsensitizers and cause minimal irritation in animal tests. In the Draize assay, three formulations, including a technical product with 100% AI and liquids containing 5 and 5.4% AI, all caused minimal (category IV) irritation.

Two products (a 0.49% aqueous formulation and a 5.4% liquid) were nonsensitizers and caused no sensitization in the Buehler assay. There were two cases associated with borates in the hand­ ler database, but there were no cases of contact dermatitis associated with direct accidental exposure (see Table 28.6). A report from Michigan documented cases of reversible alopecia in a machinist and an automotive technician, both with accidental direct exposure to the scalp with fluids containing borate (Beckett et al., 2002). The authors speculated that borate in solution, but not dry formulations of borate, posed a potential risk of toxic alopecia. Boraxassociated occupational alopecia was also reported from The Netherlands (Tan, 1970). Unlike the report from Michigan, the exposure was systemic (suspected inhalation of a washing powder containing borax) rather than topical. The most compelling evidence for a work association was reversibility of the alopecia following cessation of contact with the washing powder and demonstration of elevated urinary levels of borate. Ingestion of boric acid from a mouthwash product caused elevation of boric acid in the blood and a similar reversible alopecia in a case reported from an army hospital in San Francisco (Stein et al., 1973). CH3 C4H9O

(CH2CHO)n

CH2CHOH CH3

Butoxypolypropylene glycol

Butoxypolypropylene Glycol  Butoxypolypropylene glycol is a fly repellant used for dogs and cattle. California EPA data showed 28,875 pounds sold in California in 2007. Physical Properties  Formula (C3H6O)MULT-C4H10O; MW, 400–800; BP, 200°C; MP, 86–87°C; VP, 0.001 mm Hg at 30°C; solubility in H2O, 1 g/l at 30°C; other solubilities: soluble in kerosene and organic solvents Irritation, Sensitization, and Illness Data  A mixture containing 10% butoxypolypropylene glycol, mixed with permethrin, piperonyl butoxide, and pyrethrins (nonirritants), caused mild, transient irritation in the Draize test (category IV). The same formulation also showed sensitization in the Buehler test. There was no irritation study and no sensitization study available for the technical material. No cases involving isolated exposure to butoxypolypropylene glycol were reported in the pesticide handler database. O C

N

C2H5 C2H5

CH3 Diethyl toluamide

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

Diethyltoluamide (DEET)  DEET is an insect repellant sold in many over-the-counter formulations that range from 7 to 100% AI.3 California EPA data showed 319,649 pounds sold in California in 2007. Physical Properties  Formula, C12H17NO; MW, 191.27; BP, 160°C, 19 mm Hg; VP, 0.0056 mm Hg at 20°C; log P, 2.02; solubility in H2O, 1000 mg/l at room temperature; other solubilities: very soluble in benzene, ethyl ether, and ethanol Irritation, Sensitization, and Illness Data  Twenty-two formulations tested (7–98.3% AI) caused minimal (category IV) irritation in the Draize assay. Four formulations (25–40% AI) were nonsensitizers in the Buehler assay. Nevertheless, some formulations are reported as sensitizers on the product label because of published reports indicating that DEET can cause contact urticaria (Maibach and Johnson, 1975; von Mayenburg and Rakoski, 1983; Wantke et al., 1996). Cases of blistering irritation associated with repeated exposure to 70% DEET under humid conditions have been reported in military personnel (Lamberg and Mulrennan, 1969). Three cases listed in Table 28.6 were not classified as involving pesticide handlers and were not included in the handler database. However, two (82-1871 and 93-1422) involved apparent allergic reactions similar to those reported by Maibach (Maibach and Johnson, 1975). CH3

CH3

HN

NH N N

F3C

CH

CH

C

CH

CH

CF3

Hydramethylnon

Hydramethylnon  Hydramethylnon is an insecticide used in roach, ant, and termite baits. California EPA data showed 4310 pounds sold in California in 2007. Physical Properties  Formula, C25H24F6N; MW, 494.476; MP, 190°C; VP, 2.03  108 mm Hg, 25°C; log P, 2.31; solubility in H2O, 0.006 mg/l at 25°C; other solubilities (g/l at 20°C): acetone 360, ethanol 72, 1,2-dichloroethane 170, methanol 230, isopropanol 12, xylene 94, and chlorobenzene 390 Irritation, Sensitization, and Illness Data  Five formulations (0.365–98% AI) caused minimal (category IV) irritation in the Draize assay. The 92% technical formulation was a nonsensitizer in the Buehler test, but a mixed formulation with 0.365% hydramethylnon and 0.25% methoprene 3

Label search 2/26/2009: 536 total labels, 146 active.

751

was a weak sensitizer in the same assay. There were no cases related to hydramethylnon in the pesticide handler database. C Cl

N

N

N N

NO2

Imidacloprid

Imidacloprid  Imidacloprid is a soil, seed, or foliar insecticide formulated as a wettable powder, flowable concentrate, and as a granule. In California there were 48,189 applications in 2007, for a total of 334,623 pounds used on orchards, row crops, nursery plants, root vegetables, vineyards, corn, and cotton. Physical Properties  Formula, C9H10ClN5O2; MW, 255.69; MP, 144°C; VP, 7  1012 mm Hg, 25°C; log P, 0.57; solubility in H2O, 6.1  10  2 mg/l at 20°C; other solubilities (mg/l at 20°C): dichloromethane 6.7  104, isopropanol 2.3  104, and toluene 6.9  102 Irritation, Sensitization, and Illness Data  Twenty-five formulations of imidacloprid (ranging from 0.05% foam to 98% powdered technical material) caused minimal (category IV) irritation in the Draize assay. Three formulations (a 7.1% flowable concentrate, a 20.6% liquid, and a liquid mixture containing 12% cyfluthrin and 17% imidacloprid) caused moderate irritation in the same assay. Buehler studies on 16 formulations (0.012–76.1% AI) were negative for sensitization, as were 4 formulations (10–21.4% AI) tested in the GPMT. There was one case associated with imidacloprid in the handler database. A landscape employee helped on a pesticide application by driving the truck. He helped on landscape maintenance in between sites. At one site, the pesticide dripped from a tree onto the back of his neck, where he developed a red and slightly ulcerated rash the next day (see Table 28.6, case 2002-940). Insecticidal Soaps  Insecticidal soaps (potassium salts of fatty acids) are used to control aphids and spider mites on plants and vegetables in gardens and nurseries. Registration data indicate that skin reactions to concentrated forms of insecticidal soaps were variable. California EPA data showed 131,757 pounds used in California for 2007. Irritation, Sensitization, and Illness Data  Four products (19.5–50.9% AI) caused corrosion and two (49.2 and 51.4% AI) caused severe irritation. Four products (12.3–49.5% AI) caused moderate irritation and six products (40–49% AI) caused minimal irritation. Five dilute products (0.4–2% AI) all caused minimal irritation in the Draize assay. Buehler studies conducted on two products (0.75% insecticidal soap, 0.4% sulfur; 12.38% insecticidal soap, 6.48% sulfur) showed no evidence of sensitization. There

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was one case associated with insecticidal soap in the pesticide handler data base (see Table 28.6). CH3

N

S

N

S

O Oxythioquinox

Oxythioquinox  Oxythioquinox is a miticide, insecticide, and fungicide used for control of mite eggs and mildew on deciduous fruit. California EPA data showed minimal use in California for 2007. Physical Properties  Formula, C10H6N2OS2; MW, 234.30; MP, 171°C; VP, 2.0  107 mm Hg, 20°C; log P, 3.78; solubility in H2O, 1 mg/l at 20°C; other solubilities (g/l at 20°C): oluene 25, dichloromethane 40, hexane 1.8, isopropanol 0.9, cyclohexanone 18, dimethylformamide 10, and petroleum oils 4 Irritation, Sensitization, and Illness Data  Three products (25–92.3% AI) caused minimal (category IV) irritation in the Draize assay. The 40% flowable concentrate is a weak sensitizer in the Buehler test. Two cases associated with oxythioquinox were reported in the handler database, including one case of contact dermatitis following direct exposure (case 1982-869, Table 28.6). O CH3

C

O

O

S

CH2

C

CH

A registrant study compared the dermal sensitization potential of propargite (28.99% AI) with that of iprodione following a California fieldworker dermatitis that involved exposure to both compounds. It used a modified version of the Buehler method. Although the study had some technical deficiencies (i.e., lack of a positive control group and the use of the same animals to test both products), a number of the findings were of significance. In the range-finding portion of the study, it was determined that iprodione could be applied during the challenge tests at the maximum concentration allowed by the protocol (5%); propargite could only be applied at concentrations of 0.1%. Both materials produced less reaction during the challenge portion of the study than during the induction phase, indicating neither material was a sensitizer under the conditions tested (O’Malley et al., 1990). In contrast, technical propargite caused sensitization in the GPMT. Many cases of dermal irritation occurred followed initial registration of propargite, prior to the current electronic version of the California illness registry (Thomas and Maddy, 1975). The hazards to workers mixing and loading powdered formulations were subsequently partially mitigated by introduction of water-soluble bags. There were 105 cases associated with propargite in the handler database, occurring between 1982 and 2006. Typical cases shown in Table 28.6 (1982-1667 and 1985-1667) both involved chemical irritation following application accidents. The separate issues involved in regulating the hazards of propargite residues on crops harvested, or cultivated, with manual labor are discussed in Chapter 23. F

Propargite

Propargite  Propargite controls mites by inhibition of a mitochondrial enzyme, ATP synthase (Pridgeon et al., 2008). California use data for 2007 showed 6063 applications, for a total of 529,536 pounds used on orchards, vineyards, forage, berries, corn, cotton, and nursery plants. It contains a terminal unsaturated triple bond (an -alkyne), associated with irritation in the SICRET model and protein reactivity in the DEREK model (element IUNIQ). Physical properties  Formula, C19H26O4S; MW, 350.5; BP, decomposes at 200°C; VP, 3  107 mm Hg, 25°C; log P, 5.7; solubility in H2O, 0.215 mg/l at 20°C; other solubilities (g/l at 20°C): fully miscible with hexane, toluene, dichloromethane, methanol, and acetone Irritation, Sensitization, and Illness Data  The technical material (listed as 90.6% AI) and the liquid formulation used on cotton (73.86% AI) caused corrosion in the Draize assay. The EC (69.62% AI) caused severe irritation. Two powdered formulation (28.99 and 32% AI) nevertheless caused minimal irritation in the Draize assay.

F

F F

F F

F

F

F

F F

F F

O F

F O

F S

F NH

Sulfluramid

Sulfluramid  Sulfluramid is a sulfonamide derivative used as a cockroach and ant control bait. California EPA data for 2007 showed both limited use and limited sales. Physical Properties  Formula, C10H6F17NO2S; MW, 527.2; BP, 196°C; MP, 96°C; VP, 4.28  107 mm Hg, 25°C; log P, 6.80; solubility in H2O, insoluble at 25°C; other

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

solubilities (g/l): dichloromethane 18.6, hexane 1.4, and methanol 833 Irritation, Sensitization, and Illness Data  Three formulations (99% technical material, a 0.5% ant bait, and a roach control product with 1% sulfuramid and 0.5% chloropyrifos) caused minimal irritation in the Draize assay. A Buehler sensitization study was also negative on the roach control product (mixture of chlorpyrifos and sulfuramid described previously). There was one case associated with sulfuramid listed in the handler database (see Table 28.6) – a case of possible urticaria following accidental exposure to a termite bait.

28.2.3  Fungicides Compounds used to control fungi on plants overlap to some degree with the broader category of antimicrobial compounds. There are numerous chemical classes represented. As discussed previously, many are to one degree strong electrophiles and have the capacity to irritate or sensitize the skin. Important structural categories reviewed include phtalimido compounds, carbamates, thiocarbamates, and copper fungicides. Notable miscellaneous compounds reviewed included anilazine, chlorothalonil, chloroneb, imazalil, carboxin, flusilazole, iprodione, methylene bis(thiocyanate), and fosetyl-aluminum.

28.2.3.1  Phthalimido Compounds O

O N

S

Captan

O

CCl3

Folpet

S N

O

O

S

O

O



N

CCl3

S

CCl2 CCl2H

Captafol

N P

OCH2CH3 OCH2CH3

O Plondrel (ditalimifos)

Captan  Captan and the other phthalimido fungicides have diverse effects on fungal biochemistry, including inhibition of the synthesis of DNA and proteins, principally based on electrophilic interaction with thiols in fungal enzymes (Bushway and Hanks, 1976; Gordon, 2001). California use data for 2007 showed 6602 applications on strawberries, orchards, vineyards, and nursery crops, for a total of 449,328 pounds. Captan formulations include a wettable powder, a dust, and flowable powders. Some employ captan as the sole active ingredient, and others are combinations

753

with other fungicides and insecticides. In addition to use as a pesticide, captan has been used successfully as a treatment for pityriasis versicolor (Simeray, 1966), but it is not currently used as a human antifungal. Physical Properties  Formula, C9H8Cl3NO2S; MW, 300.59; MP, 178° C; log P, 2.8; VP, 9.0  108 mm Hg; solu­­bility in H2O, 5.1 mg/l; other solubilities (g/100 ml): chloroform 7.78, tetrachloroethane 8.15, cyclohexanone 4.96, dioxane 4.70, benzene 2.13, toluene 0.69, heptane 0.04, ethanol 0.29, and ether 0.25 at 26°C. Irritation Data  Three products (a 90% dust, a 38.52% aqueous concentrate, and an 80% granular/flake) caused minimal (category IV) irritation in the Draize assay. Sensitization Data  An 80% wettable powder caused no sensitization in the Buehler test. Data from the public domain literature suggest that captan can cause sensitization in humans. Urticaria due to captan has been documented in a gardener who reacted to captan and to captan-treated plants (Croy, 1973). Jordan and King (1977) reported a 5% sensitization rate to captan using a modified Draize test on volunteer subjects and a 10% sensitization rate on volunteers using captan in the human maximization test. Women appeared to become sensitized more frequently than men. Captan has also been reported to cause dermatitis in association with apple spraying in Scandinavia (Fregert, 1968); this has also been a relatively frequently reported problem in California. In Japan, a series of 178 patients at the Nagoya City University Medical School were routinely tested between 1977 and 1980 using the North American Contact Dermatitis Research Group’s standard allergens: 5.6% had significant positive reactions to captan. No clinical details were given in the report, but the surprisingly high percentage of patients who reacted to captan, presumably an uncommon exposure, raises the possibility that the material cross-reacts with other allergens in the standard series (Hirano and Yoshikawa, 1982). Rudner (1977) observed a similar high percentage of captan reactors in the North American Contact Dermatitis Group results in 1976 and speculated that results might be due to cross-reaction with thiurams. Illness Data  There were 14 cases associated with captan in the handler database. These included 1 case of dermatitis following direct contact (see Table 28.6). Captafol  Captafol has a chemical structure nearly identical to that of captan. It is no longer registered for agricultural use in California or elsewhere in the United States, but it had a spectrum of use similar to captan. Physical Properties  Formula, C10H9Cl4NO2S; MW, 349.1; MP, 160–161°C; log P, 3.8; VP, 8.27  109 mm Hg at 20°C; solubility in H2O, 1.4 mg/l; other solubilities: slightly soluble in most organic solvents.

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Irritation, Sensitization and Illness Data  There were no California registration data available for review, but captafol has been the topic of multiple case series and case reports in the public domain literature. Captafol accounted for 62 (28.7%) of a series of 274 cases of pesticide-associated contact dermatitis seen in Japan between 1968 and 1970 (Matsushita et al., 1980). In a similar series, 22 (18.2%) of 121 Korean farmers likewise reacted to the material (Lee et al., 1981). Cottel (1972) observed several cases of San Joaquin Valley orchard farmers with positive patch test responses to a 0.1% aqueous preparation of captafol. Irritant and allergic contact dermatitis was also seen in 23% of 133 New Zealand timber workers tested with the material by Stoke (1979). Camarasa (1975) found 4 of 7 ill workers from a captafol packaging plant had 3 patch test responses to 1% captafol. An outbreak of dermatitis due to captafol sensitivity was also seen among a group of 36 workers on a Kenyan coffee plantation (Verhagen, 1974). Urticaria and asthma were part of the clinical picture reported, affecting 7 (17.1%) of 41 workers in a captafol packing operation in a chemical shed (Camarasa, 1975). The similar occurrence of asthma and contact dermatitis in a welder employed by a maintenance firm that serviced captafol distribution plants was reported by Groundwater (1977). Thus, captafol is apparently capable of causing both delayed and immediate types of hypersensitivity, as well as irritant dermatitis. There was one case associated with captafol in the handler database, but it did not involve dermatitis following accidental direct contact. Folpet  Folpet is a fungicide previously used for protection of fruits, berries, and ornamentals. It currently has limited use in agriculture but some remaining use as an industrial preservative. Physical Properties  Formula, C9H4Cl3NO2S; MW, 296.56; MP, 177°C; log P, 2.85; VP, 1.58  10-5 mm Hg; solubility in H2O, 1 mg/l at 20°C; other solubilities (g/l): chloroform 87, benzene 22, and isopropanol 12.5 Irritation Data  The 88% technical material (a solid formulation with distilled water) and a dilute mixture with 0.5% folpet and 0.5% bis(tributyltin) oxide caused minimal skin irritation in the Draize test, but the 49.4% liquid folpet concentrate labeled for use as a paint preservative caused moderate irritation. Sensitization Data  Technical folpet (an 88% solid) is a sensitizer in the guinea pig maximization assay. Illness Data  There were no cases associated with its use in the handler database. Plondrel  Plondrel (ditalimifos) structurally resembles captafol and captan, but it can also be considered an

organophosphate because it has a side chain containing a phosphothioate group. It is not currently registered in the United States. Physical Properties  Formula, C12H14NO4PS; MW, 299.284; log P, 3.48; VP, 2.38  108 mm Hg; solubility in H2O, 133 mg/l; other solubilities: NA Irritation, Sensitization, and Illness Data  No registration data were available for review. Reports in the public domain literature included an episode reported in 1975 by van Ketel involving four workers spraying ditalimifos on roses who subsequently developed dermatitis. All four reacted to 0.1% Plondrel in petrolatum, but no reactions occurred in 20 control subjects tested with the same material (van Ketel, 1975). Van Ketel subsequently reported a third case of hand eczema in a 21-year-old florist who had a 3 reaction to 0.1% plondrel (van Ketel, 1977).

28.2.3.2  Carbamates

N

  N C

NHCO2CH3 NHC4H9

O Benomyl

Benomyl  Benomyl is a benzimidazole compound with a carbamate moiety but has no activity as a cholinesterase inhibitor. It is used in the control of many diseases of fruits, nuts, vegetables, and ornamental plants. Most of the available data on its dermal effects derive from the public domain literature. Physical Properties  Formula, C14H18N4O3; MW, 290.32; MP, 140°C; log P, 2.12; VP, 3.7  109 mm Hg; solubility in H2O, 2–3.8 mg/l; other solubilities (g/kg): acetone 18, chloroform 94, dimethylformamide 53, ethanol 4, heptane 0.4, and xylene 10 Irritation and Sensitization Data  Guinea pig tests of benomyl for irritancy conducted by the manufacturer at 12.5 and 25% aqueous dilutions were reported to be negative (Matsushita and Aoyama, 1981). However, the maximization test conducted in the same study showed 2% benomyl to be a potent experimental allergen. The first report implicating benomyl as a contact allergen appeared in 1972. Seven Japanese women employed in a greenhouse by a carnation grower developed dermatitis of exposed skin after benomyl was sprayed there on two occasions. No cases occurred until 2 weeks after the second spraying. The seven patients had 2 reactions to a 1:10 dilution of benomyl in olive oil; three control subjects were ­negative

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

(Savitt, 1972). Van Ketel (1976) also reported a case of benomyl sensitivity, confirmed by patch testing (with a 1% preparation that elicited no reaction from 10 controls), in a begonia grower. A second report from The Netherlands also highlighted the occurrence of benomyl hypersensitivity in nursery workers and florists (van Joost et al., 1983). The preceding cases illustrate the capacity of foliar residues of benomyl to cause allergic contact dermatitis in nursery workers. Zweig et al. (1983) demonstrated that exposure up to 5.4 mg/person-hour to benomyl is also a potential problem in strawberry harvesting. Everhart and Holt (1982) studied benomyl applicators and noted a maximum total exposure of 26 mg of benomyl in mixing/loading operations and markedly lower total exposures associated with field residue exposure (12 mg) and home use of the material. Hargreave (1983) noted the possibility of exposure from handling treated commodities. He demonstrated persistent benomyl residues on litchi nuts up to 15 days after postharvest treatment in a dipping process: 20 ppm of benomyl in the skin and 1.3 ppm in the flesh of the nut. There were 13 cases associated with benomyl in the pesticide handler database. Case 1985-448 describes a case of contact dermatitis following direct accidental exposure (see Table 28.6).

755

S

S CH2

NH C S

CH2

NH C S

CH2

Mn/Zn

NH C S

CH2 NH C S

Mancozeb S

Mn N N>1

S

Maneb S CH2

NH C

S

CH2 NH C S

Zn N N>1

S Zineb

CH3

CH3 CH3

N C S S

Zn

S C N S

Ziram NH

CH3

S NH CH3

CH3 CH3

N C S

S C N

S

S

CH3

Thiram

S

O

NH C NH C OCH3 NH C NH C OCH3 S O Thiophanate methyl

Thiophanate Methyl  Thiophanate methyl is a systemic fungicide, usually applied as a dust or powder, used on vegetables, beans, nuts, potatoes, and turf. Physical properties  Formula, C12H14N4O4S2; MW, 342.40; MP, 172°C; log P, 1.40; VP, 7.13  108 mm Hg; solubility in H2O, 26.6 mg/l; other solubilities (g/kg): acetone 58.1, cyclohexanone 43, methanol 29.2, acetonitrile 24.4, and ethyl acetate 11.9 Irritation and Sensitization Data  Eleven formulations (1.5– 96.2% thiophanate) caused minimal irritation in the Draize assay. Two formulations tested in the Buehler assay (an 85% granule and a mixture of 28.5% thiophanate and 51.42% flutolanil) proved negative for sensitization. However, the 96.2% dust formulation caused sensitization in the GPMT. Illness Data  There was one case associated with thiophanate methyl in the handler database, but it did not involve dermatitis following accidental direct contact.

28.2.3.3  Thiocarbamates The thiocarbamate group of fungicides structurally resembles the rubber accelerator disulfiram (Antabuse, tetraethylthiuram disulfide; CAS No. 97-77-8), a common sensitizer present in both the European and the North American standard patch test series (Adams and Fischer, 1990). Cellular toxicity of the compounds depends on the oxidant effects related to the thiocarbamate disulfide bridge and reducing effects of the SH groups (Grosicka et al., 2005). Targets of nucleophilic inhibition include enzymes involved in ATP production, the Krebs cycle, and conversion of glucose to pyruvate and fatty acids to acetyl coenzyme A (Hurt et al., 2001). The thiocarbamates may also contain the sensitizer ethylene thiourea (ETU) as a contaminant (Bruze and Fregert, 1983; Hajslová et al., 1986; Hwang et al., 2001; Meding et al., 1990), and most yield ETU as a metabolic product following occupational exposure (Sciarra et al., 1994; Swaen et al., 2008). ETU is an electrophile rather than a nucleophile (U.S. EPA, 2001). Thiram, Ziram, Zineb, Maneb, and Mancozeb  The prototype thiocarbamate fungicide, thiram (thiuram), is simply the methyl analog of disulfiram, and experimentally it has a similar effect on the metabolism of alcohol (Freundt and Netz, 1977). The structure of ziram is very similar to that of thiram, but the compound contains a zinc atom between the two

756

atoms of sulfur. It is also similar to zineb, which is a zinc/ thiocarbamate polymer, and to the manganese/thiocarbamate polymer maneb. Mancozeb is a related product containing both zinc and manganese. Physical Properties Thiram: Formula, C6H12N2S4; MW, 240.44; MP, 155– 156°C; BP, 129°C (at 20 mm Hg); log P, 1.73; VP, 1.73E-05 mm Hg; solubility in H2O, 30 mg/l; other solubilities (g/l at 25°C): ethanol 10, acetone 80, chloroform 230, and hexane 0.04; (g/l at 20°C) dichloromethane 170, toluene 18, and isopropanol 0.7 Ziram: Formula, C6H12N2S4ZN; MW, 305.83; MP, 246°C; log P, 1.23; VP, 7.5109mm Hg at 0°C; solubility in H2O, 65 mg/l; other solubilities (g/100 ml at 25°C): ethanol 0.2, acetone 0.5, benzene, 0.5, carbon tetrachloride, 0.2, ether 0.2, and naphtha 0.5 Zineb: Formula, C4H6N2S4Zn; MW, 275.75; MP, decomposes at 157°C without melting; log P, 1.3; VP, 7.5  109 mm Hg at 0°C; solubility in H2O, 10 mg/l; other solubilities: soluble in carbon disulfide, pyridine, benzene, and chloroform Maneb: Formula, C4H6MnN2S4; MW, 275.75; MP, decomposes without melting at 157°C; log P, 0.620; VP, 1.0  107 mm Hg; solubility in H2O, 10 mg/l; other solubilities: soluble in carbon disulfide, pyridine, benzene, and chloroform Irritation Studies  A 38.8% maneb liquid caused moderate (category III) irritation in the Draize assay. A mixture of 22.1% copper sulfate and 30.4% mancozeb caused moderate irritation in the Draize assay. The remaining thiocarbamate products caused minimal irritation. These include four thiram products (technical powder with 98.8% AI, a 77% granule, and two mixtures with 3.07–4.2% AI), a 51% slurry of ziram, a mixed powder with 8% maneb and 0.01% streptomycin sulfate, and six mancozeb products (with 15–80% AI). Sensitization Studies  A 77% granular thiram product caused sensitization in the Buehler assay. Thiram also caused sensitization in the LLNA at concentrations of 1% or higher (de Jong et al., 2002), and there have been numerous reported cases of sensitization in the clinical literature (Cronin, 1980; Schultz and Hermann, 1958; Shelley, 1964). A 47.8% liquid and a mixture of 63% mancozeb and 15% thiophanate methyl caused sensitization in the Buehler assay, but two mancozeb products (33.9% liquid and 82.3% powder) were negative in the same assay. Mancozeb technical powder (exact percentage AI not reported) caused sensitization in the GPMT, as did a mixture of 30.4% mancozeb and 22.1% copper sulfate. Matsushita tested maneb and zineb experimentally with the guinea pig maximization procedure and found both compounds to be potent sensitizers with a high degree of

Hayes’ Handbook of Pesticide Toxicology

mutual cross-reactivity. Concentrations of 5% or more were found to be irritating (Matsushita et al., 1976). Cases of allergic reactions to maneb documented with provocation (patch) testing have been reported in the clinical literature. Typical cases described from The Netherlands included two office workers who had purchased maneb spray to care for the plants in their office and a 51-year-old woman who worked as an assistant in a flower shop (Nater et al., 1979). Similar cases have been reported from the United States (Adams and Manchester, 1982), Italy (Peluso et al., 1991), and Germany (Koch, 1996). A case reported from Japan identified mancozeb as a cause of allergic contact dermatitis and photodermatitis (Higo et al., 1996). Cases in the handler database  There were 18 cases associated with thiocarbamates in the handler database. These included 4 episodes of direct accidental exposure to ziram and a case of facial dermatitis associated with thiram dust trapped underneath the respirator (see Table 28.6). Cases of suspected allergy included a possible allergic reaction after spraying a maneb-containing formulation of Dithane (1984-811) and a cervical rash in an applicator with a history of sensitivity to mancozeb (1986-619).

28.2.3.4  Copper Fungicides Copper compounds are used as both fungicides and antimicrobial agents. Copper has been identified as a sensitizer in the public domain literature based on human case reports (Rademaker, 1998; Verhagen, 1974), and most of the copper fungicides are labeled as potential sensitizers even where there are negative animal sensitization studies (e.g., copper hydroxide). Data on representative compounds, cuprous oxide, cupric oxide, and copper naphthenate are reviewed here. There were 19 cases associated with copper fungicides in the handler database, most frequently involving copper hydroxide and copper sulfate. These included numerous cases of dermatitis following direct contact (see Table 28.6). Copper (II) Hydroxide  California use data for 2007 showed 36,894 applications, for a total of 2,220,953 pounds used on orchards, cotton, vineyards, row crops, and nursery crops and for landscape maintenance. Physical Properties  Formula, CuH2O2; MW, 97.56; MP, decomposes, with loss of water; solubility: in water, 2.9 mg/ l at 25°C, pH 7 Irritation Data  A liquid product containing 14.77% copper hydroxide and a 3.6% EC caused moderate irritation in the Draize assay. Seven solid products (wettable powders, dry flowables, granules, 40.87–90% AI) caused minimal irritation, as did five liquid products (3.1–77% AI). Sensitization Data  A 3.6% liquid caused no sensitization in the Buehler test; a 40.87% granular product was negative in the GPMT.

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Copper (I) Oxide, Cuprous Oxide  California agricultural use data for 2007 showed 5396 applications, for a total of 263,679 pounds used for orchards, vineyards, row crops, and nurseries.

20 dynes/cm  0.020 N/m at 20°C; VP, 133 mPa (0.001 mm Hg); solubilities: practically insoluble in water; soluble in most organic solvents; moderately soluble in petroleum oils

Physical Properties  Formula, Cu2O; MW, 143.09; BP, 1800°C; MP, 1235°C; VP, negligible; solubilities: soluble in ammonium hydroxide; in hydrochloric acid forming copper(I) chloride, which dissolves in excess hydrochloric acid; with dilute sulfuric acid or dilute nitric acid the cupric salt is formed and half copper is precipitated as the metal; practically insoluble in organic solvents; soluble in dilute mineral acids; soln of ammonia and its salts; insoluble in water

Irritation Data  A gel paste containing 40% borax and 18.16% copper naphthenate gel paste caused corrosion in the Draize assay. An 8% liquid caused moderate irritation. Liquid products containing 2 and 8% AI and an 80% EC caused severe irritation. Copper naphthenate was also associated with one case (1987-1724) described as a chemical burn.

Irritation Data  A liquid product containing 40.41% copper (I) oxide (cuprous oxide) and 3.8% zinc 2-pyridinethiol-1-oxide caused corrosion in the Draize assay. A liquid mixture of 24.59% copper (I) oxide and 1.86% tributyltinoxide and a 46.5% liquid caused severe irritation. Four liquid products (43–66.9% AI) and an 82% dust caused moderate irritation. Although the previous data suggest a consistent pattern of irritation associated with copper (I) oxide, 95% copper (I) oxide powder, and 11 liquid products (principally paint) containing 40.36–65% AI, caused minimal irritation.

28.2.3.5  Fungicides with Miscellaneous Structures

Sensitization Data  A mixture of 37% copper (I) oxide and 1.86% 4,5-dichloro-2-n-octyl-3(2 H)-isothiazolone and a mixture of 40.41% copper (I) oxide and 3.8% zinc 2pyridinethiol-1-oxide caused no sensitization in the Buehler test. Cupric Oxide, Copper (II) Oxide  A total of 372,399 pounds were reported used in California agriculture in 2007, principally for treated lumber. Physical Properties  Formula, CuO; MW, 79.55; BP, 1026°C; MP, 1326°C; solubility: practically insoluble in water and alcohol; soluble in dilute acids, alkali cyanides, and ammonium carbonate solution; slowly soluble in ammonia Irritation Data  A mixture of 14.07% cupric oxide and 35.46% chronic acid caused corrosion in the Draize assay. A 97.6% copper (II) oxide powder caused moderate irritation. Sensitization Data  Cupric oxide technical material caused no sensitization in the Buehler assay. Copper Naphthenate  Copper napthenate is used as a preservative, treating roof shingles, fences, and other wooden products. California data showed 1,009,571.73 pounds sold in California in 2007. Physical Properties  Copper in mixture of cyclopentyl and cyclohexyl carboxylic acids, with MW of 120–700: MW, 405.86 (variable); BP, 154.4–201.7°C; surface tension,

Cl Cl

N N

NH N

Cl

Anilazine

Anilazine  Anilazine is a foliar and turf fungicide that has not been registered in California since 1990 but is currently being used elsewhere in the United States. Physical Properties  Formula, C9H5Cl3N4; MW, 275.52; MP, 160°C; log P, 3.88; VP, 6.2  109 mm Hg; solubility in H2O, 8 mg/l at 30°C; other solubilities (g/100 ml at 30°C): toluene 5, xylene 4, and acetone 10 Irritation and Sensitization Data  Data on dermal irritation and sensitization from animal studies were not available for review, but the product has been reported as a human sensitizer in tomato harvesters (Schuman and Dobson, 1985; Schuman et al., 1980) and in lawn care workers (Mathias, 1997). Illness Data  The sample case from the handler database described in Table 28.6 (1992-732) was suspected to be caused by an allergic reaction, but patching testing was not carried out. O

CH3

S

C NH O Carboxin

Carboxin  Carboxin is a systemic fungicide and seed protectant. For 2007, California agricultural use data showed 1256 pounds applied to stored corn, onions, cotton, and other commodities.

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Physical Properties  Formula, C12H13NO2S; MW, 235.3; MP, 94°C; log P, 2.14; VP, 1.80  107 mm Hg; solubility in H2O, 199 mg/l; other solubilities (g/l at 20°C): acetone 221.2, methanol 89.33, and ethyl acetate 107.7 Irritation and Sensitization Data  It is a nonirritant in the Draize test. A mixture containing carboxin (15%), PCNB (15%), and metalaxyl (3.12%) was a nonsensitizer in the Buehler assay. Cl OCH3 CH3O Cl Chloroneb

Chloroneb  Chloroneb is a fungicide used for control of seedling diseases in beans, cotton, and soybeans. For 2007, California agricultural use data showed 1256 pounds applied to stored corn, onions, cotton, and other commodities. Physical Properties  Formula, C8H8Cl2O2; MW, 207.06; MP, 133–135°C; BP, 268°C; log P, 3.44; VP, 3  103 mm Hg; solubility in H2O, 8 mg/l; other solubilities (g/kg at 25°C): acetone 115, xylene 89, dimethylformamide 118, and dichloromethane 133 Irritation, Sensitization, and Illness Data  A formulation containing 30% chloroneb and 3.5% metalaxyl caused moderate irritation in the Draize test. The same mixture was a nonsensitizer in the Buehler assay. There were no cases associated with either chloroneb or carboxin in the handler database. C

N

Cl

Cl

Cl

C

N

Cl Chlorothalonil

Chlorothalonil  Chlorothalonil is an electrophile that inhibits thiol enzymes important for fungal spore germination (Leroux et al., 2002) and sulfhydryl groups important in glycolysis and fungal respiration. Depending on species and route of administration, various glutathione conjugates can be measured as urinary metabolites (Parsons, 2001). It has a soil half-life of approximately 2 months and is stable on exposure to ultraviolet light. For 2007, California use data showed 14,852 applications, for a total of 734,604 pounds used on grains, orchards, row crops, vineyards, and nurseries and for landscape maintenance.

Physical properties  Formula, C8Cl4N2; MW, 265.9; MP, 250°C; BP, 350°C; log P, 3.05; VP, 5.7  107 mm Hg; solubility in H2O, 0.81 mg/l; other solubilities (g/kg at 25°C): acetone and dimethyl sulfoxide 20, cyclohexanone and dimethylformamide 30, kerosene 10, and xylene 80 Irritation data  A liquid mixture of 20.06% chlorothalonil and 10.1% 3-iodo-2-propynyl butylcarbamate caused corrosion in the Draize assay. A 50% chlorothalonil liquid and a mixture of 14.7% methylene bis(thiocyanate) and 14.5% chlorothalonil caused severe irritation. Four solid products (82–98.6% AI) and three liquid products (40.8% flowable concentrate, a mixture of 39.3% chlorothalonil and 2.13% triadimefon, and 12.5% chlorothalonil) caused moderate irritation in the Draize assay. Three solid (50–75% AI) and seven liquid products (0.333%  54.9% AI) caused minimal irritation. A summary of manufacturer conducted studies (Parsons, 2001) states that chlorothalonil causes skin irritation principally after cumulative exposure. Public domain literature shows that 0.1% chlorothalonil in acetone is a moderate cutaneous irritant (irritation score  2.71) in experimental studies with New Zealand white rabbits; 0.1% chlorothalonil in petrolatum is much less irritating (irritation score  0.71), and 0.1% chlorothalonil in saline (irritation score  0.04) is nonirritating (Flannigan and Tucker, 1985b; Flannigan et al., 1986). An outbreak of irritant dermatitis and conjunctivitis in a Portuguese tent manufacturing operation that used chlorothalonil-impregnated fabric has been described (Lensen et al., 2007). Of 11 workers employed in the operation, 3 resigned prior to the investigation. The remaining 8 workers reported erythema, pruritus, and scaling of the eyelids, face, and arms; conjunctivitis; and pharyngitis after handling batches of treated fabric. Symptoms typically disappeared during time away from work. Patch testing with the fabric showed typical irritant responses; testing with the standard allergen series showed some positive responses (to nickel and thimerosol) judged not relevant to the workplace illness cluster. Tests with a textile allergen series and with 0.01% chlorothalonil (tested separately in petrolatum and saline) were negative. The authors considered cumulative irritation from airborne exposure to chlorothalonil to be the cause of the outbreak. Sensitization data  A liquid product containing 40.4% chlorothalonil caused sensitization in the Buehler test; a 54% liquid product was negative for sensitization. Other animal sensitization studies conducted by the manufacturer have been summarized by Parsons (2001), who concluded that chlorothalonil is a weak sensitizer. Of the 10 studies conducted with either the Buehler or maximization methods, 7 showed some degree of sensitization (Parsons, 2001).

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

A public domain study (Boman et al., 2000) using reagent-grade chlorothalonil (99% purity) showed chlorothalonil to be a sensitizer in both the cumulative contact enhancement test (a variant of the GPMT) and the LLNA. The EC3 in the LLNA ranged from 0.002 to 0.035% w/v, depending on experimental technique, comparable to EC3 values obtained with DNCB and oxazolone in the same laboratory. Contact dermatitis has been reported in vegetable growers, woodworkers, and flower growers (Bruynzeel and van Ketel, 1986). Patch testing is performed with 0.01% chlorothalonil in petrolatum. This is a marginally irritant concentration when butanol is used as the patch test vehicle (O’Malley et al., 1995). Penagos et al. (1996) reported chlorothalonil as a possible cause of erythema dyschromicum perstans (ashy dermatitis). Positive patch test reactions to chlorothalonil (0.001% in acetone) were observed in 34 of 39 banana farm workers presented with erythema dyschromicum perstans-like dermatitis. Biopsies from all patients were compatible with a chronic pigmented dermatitis or erythemadyschromicum perstans-like dermatitis (Penagos, 2002; Penagos et al., 1996). Cases reported from Japan include a case of photoallergic contact dermatitis (Matsushita et al., 1996) and a case of contact allergy (Matsushita, 1995). Related cases of chlorothalonil skin allergy were accompanied by cases of asthma (Huang et al., 1995). Asthma was a primary endpoint in an employee of a fungicide formulating operation reported from England. There was no history of skin reactions, but an inhalation challenge with 12.5 g chlorothalonil in 250 g lactose mixture (for 30 min) resulted in a 20% decline in forced expiratory volume beginning 3 h after exposure. It was not possible to confirm that the mechanism was allergic rather than allergic by testing a specific chlorothalonil IgE by RAST, by skin testing for either delayed or immediate allergy, or by testing controls using the same inhalation challenge administered to the patient. Chlorothalonil is used as a wood preservative in northern Europe. Johnsson et al. (1983) reported an epidemic of contact dermatitis in a Norwegian woodenware factory. Fourteen out of 20 workers had work-related skin complaints, and 7 workers had contact dermatitis. Bach and Pedersen (1980) reported contact dermatitis in a cabinet maker in contact with a chlorothalonil-containing wood preservative. Three similar cases of contact dermatitis related to a chlorothalonil wood preservative have also been reported from Germany (Spindeldreier and Deichmann, 1980). Fatal toxic epidermal necrolysis has been attributed to chlorothalonil. Lord et al. (1984) reported on a 30-year-old navy pilot who had played 81 holes of golf during the week prior to developing toxic epidermal necrolysis. The golf course had been sprayed with chlorothalonil. The authors stated that special photographic techniques using ultraviolet light demonstrated chlorothalonil on the deceased’s golf

759

clubs, balls, and shoes. In a case reported by Dannaker et al. (1993), chlorothalonil was associated with contact urticaria in a nursery worker and an anaphylactoid reaction on provocation testing with 1% aqueous chlorothalonil. The capability of chlorothalonil to cause allergy was contested by the manufacturer during the 1990s (Eilrich and Chelskey 1991). Product labels currently available online at http://www.cdms.net for a mixture of 72% chlorothalonil and 4.4% metalaxyl, an 82.5% granule, and a mixture of 33.1% chlorothalonil and 3.3% state in the note to physicians that “temporary allergic skin reactions may respond to treatment with oral antihistamines and topical or oral steroids.” The precautionary statements also state that “prolonged or frequently repeated skin contact may cause allergic reaction in some individuals.” California Illness Data  There were 18 cases associated with chlorothalonil in the handler database, including 3 that followed accidental direct exposure. All appeared to be cases of irritation based on the short intervals between exposure and onset of the rash (see Table 28.6). The case of sensitization reported by Dannaker et al. (1993) was related to chlorothalonil residue exposure and not included in the handler cases. Cl OH C

N

N

Fenarimol

Fenarimol  Fenarimol is a pyrimidine ergosterol biosynthesis inhibitor used as a fungicide (Proenca et al., 2003). It also structurally resembles the organochlorine insecticides DDT and dicofol, and it is considered a suspect endocrine disruptor (de Castro et al., 2007). For 2007, California use data show 3748 applications, for a total of 4386 pounds used principally on orchard, vineyard, and nursery crops. Physical properties  Formula, C17H12Cl2N2O; MW, 331.20; MP, 118°C; log P, 3.6; VP, 2.25  107 mm Hg; solubility in H2O, 14 mg/l; other solubilities: soluble in acetone, acetonitrite, benzene, chloroform, and methanol, but only slightly soluble in hexane Irritation, Sensitization, and California Illness Data  Granules containing 0.78% fenarimol caused minimal irritation in the Draize assay. There were no dermal sensitization studies available for review. There were three cases associated with fenarimol in the handler database, but there were no cases that occurred following direct exposure (see Table 28.6).

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760

N

F N

CH2 F

Si

Flusilazole

CH3

CH O CH2

CH2

Cl

N

N

CH

Imazalil N Cl

Flusilazole  Flusilazole is an organosilicon compound that functions as an inhibitor of ergosterol biosynthesis at the 14-demethylase enzyme step (Henry, 1989). It is formulated as dry granules and as an EC, and it is used for control of ascomycetes and other fungi on cereals, fruits, and vegetables. It is not currently registered in the United States. Physical properties  Formula, C16H15F2N3Si; MW, 315.40; MP, 54°C; log P, 3.7; VP, 2.93  107 mm Hg; solubility in H2O, 54 mg/l; other solubilities: soluble in many organic solvents Irritation, sensitization, and California Illness Data  A 60.7% formulation caused moderate irritation in the Draize test, but a 20% formulation was a nonsensitizer in the Buehler assay. There were no cases associated with flusilazole in the handler database. O C2H5

O

P

+ O- Al3 3

Fosetyl Al

Fosetyl-Aluminum  Fosetyl-aluminum is an aluminum salt of ethyl phosphonic acid that disrupts plant amino acid synthesis. It is active against Oomycetes, Alternaria, and Penicillium. For 2007, California use data showed 13,292 applications, for a total of 293,616.5586 pounds used on orchards, row crops, nursery crops, and strawberries. Physical Properties  Formula, C6H18AlO9P3; MW, 354.10; MP, 215°C; log P, 2.1 to 2.7; VP, 7.5  1010 mm Hg, 25°C mm Hg; solubility in H2O, 1.20 E05 mg/l; other solubilities: practically insoluble in acetonitrile and propylene glycol (80 mg/l) Irritation, Sensitization, and California Illness Data  A product with 80% wettable powder and a 70.2% granular product caused minimal irritation in the Draize assay. The 80% powder caused no sensitization in the Buehler test. There were no cases associated with its use in the handler database.

Imazalil  Imazalil (enilconazole) is a systemic fungicide active against benzimidazole-resistant strains of fungi, an electrophilic inhibitor of ergosterol biosynthesis at the 14-demethylase enzyme step (Stenersen, 2004). 2007 California use data showed 14,421 pounds applied on citrus and nursery crops. Physical properties  Formula, C14H14Cl2N2O; MW, 297.2; MP, 52.7°C; BP, 347°C; log P, 3.82; VP, 1.2  106 mm Hg; solubility in H2O, 180 mg/l; other solubilities: etone, dichloromethane, ethanol, methanol, isopropanol, xylene, toluene, and benzene 500 g/l at 20°C Irritation, Sensitization, and California Illness Data  A 31% flowable concentrate caused severe irritation in the Draize assay. Technical solid (98.5% AI) and four liquid products (50% AI, 13.8% AI, 10% AI, and a mixture of 2% imazalil and 1% proprioconazole) caused minimal irritation. Technical imazalil caused no sensitization in the GPMT and the Buehler test was negative on a 13.5% liquid. Cases of sensitivity to the compound have been prev­ iously described in Europe, associated with veterinary use (van Hecke and de Vos, 1983) and in banana production in Central America (Penagos, 1993). There were no cases associated with imazalil in the handler database. CH(CH3)2 O

NH O

C

N

Cl

N O

Cl

Iprodione

Iprodione  Iprodione induces osmotic sensitivity in fungi by inhibiting enzymes involved in cell wall production (Cui et al., 2002). 2007 California use data showed 16,703 applications, for a total of 251,168 pounds used on grains, berries, row crops, orchards, and nursery crops. Physical Properties  Formula, C13H13Cl2N3O3; MW, 330.17; MP, 136°C; log P, 3.00; VP, 3.75  109 mm Hg; solubility in H2O, 13.9 mg/l; other solubilities (g/l at 20°C): ethanol 25,

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

methanol 25, acetone 300, dichloromethane 500, and dimethylformamide 500 Irritation and Sensitization Data  A product containing 75% granules caused moderate irritation in the Draize assay. Three liquid products (23% AI, 28.6% AI, and a mixture with 19% iprodione/20.4% thiophanate-methyl) and a 50% granular product caused minimal irritation. The 28.6% liquid, the mixture with thiophanate, and the 75% granular product were nonsensitizers in the Buehler test. California Illness Data  There were five cases associated with iprodione in the handler database, including three cases of dermatitis that followed direct contact (see Table 28.5). NCS CH2 SCN Methylene bisthiocyanate

Methylene Bis(thiocyanate)  Methylene bis(thiocyanate) (MBTC) is a reactive nucleophile, inhibiting respiration by binding with fungal cytochromes. Sales data showed 22,896 pounds of MBTC sold in California during 2007. Registered products are intended for use as an industrial biocide in water-cooling systems, for use in pulp and paper mill operations, and as a wood preservative. Physical Properties  Formula, C3H2N2S2; MW, 130.19; MP, 105–107°C; log P, 0.620; VP, 1.97E-03 mm Hg; solubility in H2O, 2.72E04 mg/l; other solubilities (at 20°C): 100 mg/ml in DMSO, 10–50 mg/ml in 95% ethanol, 100 mg/ml in acetone Irritation, Sensitization, and California Illness Data  A liquid product containing 10% MBTC caused corrosion in the Draize assay. A similar product that also contained 10% MBTC was reported to cause moderate irritation. A mixture containing 0.2% MBTC, 0.1% chlorpyrifos, and 0.2% 2-(thiocyanomethyl thio)benzothiazole (TCMTB) caused minimal irritation. The only sensitization study reviewed involved the mixture of MBTC, chlorpyrifos, and TCMTB and did not show any evidence of an allergic response. There were no cases associated with MBTC in the hand­ ler database. CN Cl

N

C CH2

N N CH2CH2CH2CH3

Myclobutanil

Myclobutanil  Myclobutanil inhibits cytochrome P450 (CYP) 51 activity and biosynthesis of ergosterol by fungi (Tang et al., 2005). For 2007, California agricultural use data showed 19,474 applications, for a total of 65,161 pounds used on orchards, cotton, row crops, vineyards, and nursery crops and for landscape maintenance.

761

Physical Properties  Formula, C15H17ClN4; MW, 288.78; MP, 63–68°C; BP, 202–208°C at 1 mm Hg; log P, 2.94; VP, 1.6  106 mm Hg; solubility in H2O, 142 mg/l; other solubilities: soluble in common organic solvents such as ketones, esters, alcohols, and aromatic hydrocarbons – all 50–100 g/l; insoluble in aliphatic hydrocarbons Irritation Data  Products containing 25 and 27% emulsifiable concentrates caused severe irritation in the Draize assay. A 19.7% liquid caused moderate irritation. Three liquid products (1.55–84.5% AI) and two solid products (0.39% powder and 0.62% granules) caused minimal irritation. Sensitization Data  Two products (1.55% myclobutanil and 2.5% permethrin concentrate and a 6% liquid) caused no sensitization in the Buehler test. A 20% liquid was negative in the GPMT. California Illness Data  There were four cases associated with myclobutanil in the handler database, including two associated with direct exposure (see Table 28.6). Cl Cl

Cl

Cl

Cl NO2 PCNB

Pentachloronitrobenzene (PCNB)  Pentachloronitroben zene (PCNB) has broad antifungal activity. The mode of action is probably similar to that of chlorothalonil, inhibiting fungal enzymes (e.g., cytochrome reductase) by electrophilic interaction with thiol groups (Fall and Murphy, 1984; Hall, 2000). California agricultural use data for 2007 showed 1302 applications, for a total of 30,663 pounds used on broccoli, Brussels sprouts, cotton, nursery crops, and turf. Physical properties  Formula, C6Cl5NO2; MW, 295.34; MP, 144°C; BP, 328°C at 760 mm Hg with some decomposition; log P, 4.22; VP, 5  105 mm Hg at 20°C; solubility in H2O, 0.44 mg/l; other solubilities (g/l at 20°C): toluene 1140, methanol 20, and heptane 20 Irritation and Sensitization Data  A 24.3% liquid caused severe irritation in the Draize assay and a 15% granular product caused moderate irritation. A 40% liquid and two solid products (95% technical and a dust containing 25% PCNB and 6.25% metalaxyl) caused minimal irritation. Three products (15% granular, 23.8% EC, and 95% technical material negative) were negative for ­sensitization

762

in the Buehler test. However, two patch tests indicating sensitization to PCNB were reported in 2 of 39 nursery workers surveyed in California. It was not possible to determine whether the reactions were relevant to a workplace exposure or represented de novo sensitization (O’Malley and Rodriguez, 1998a,b; O’Malley et al., 1995). Positive reactions were also reported in a survey of farmers from Japan (Kambe et al., 1976). Twelve cases related to PCNB were also reported based on surveillance data from Japan, but no patch test information or other clinical details were available (Horiuchi et al., 2008). California Illness Data  There was one case following direct contact with PCNB reported in the handler database (see Table 28.6). Elemental Sulfur and Sulfur dioxide  Sulfur is a broadspectrum fungicide. Its mode of action remains unknown but probably involves transformation products rather than elemental sulfur per se. Sulfur dioxide, for example, interacts with multiple cellular enzymes and is used to control fungal pathogens in commodities. For 2007, California agricultural use data showed 126,711 applications, for a total of 46,056,219 pounds used on grains, orchards, row crops, vineyards, nursery crops, pastureland, and berries. Physical properties Elemental Sulfur  Formula, S; MW, 32.06; MP, 75.5°C; BP, 10.05°C; log P, not available; VP, 3  103 mm Hg; solubility in H2O, insoluble in water; other solubilities: 1 g/2 ml carbon disulfide Physical properties Sulfur Dioxide  Formula, SO2; MW, 64.065; MP, 112.8–120°C; BP, 444.6°C; log P, not available; VP, 3.95  106 mm Hg; solubility in H2O, 8.5% at 25°C; other solubilities: soluble in chloroform, ether, acetic acid, and sulfuric acid Irritation data  Two sulfur products (a granular mixture of 80% sulfur and 1.56% imidacloprid and a liquid mixture of 10% sulfur and 0.25% pyrethrins) caused moderate irritation in the Draize assay. Seventeen products (0.2–99% AI) caused minimal irritation. No data are available on sulfur dioxide because of its physical properties. California Illness Data  There were 68 cases associated with elemental sulfur in the handler database. These included 11 cases of dermatitis following direct contact. Several cases of interest are included in Table 28.6. Case 1986-968 suggested the possible effect of perspiration in increasing the effect of skin contact with sulfur dust. There were two separate episodes of spontaneous ignition during application (1997-1117 and 2000-485, 486); similar episodes during aerial application are discussed in Chapter 88. All of the cases cited previously are consistent with an irritant mechanism.

Hayes’ Handbook of Pesticide Toxicology

Three cases of irritation related to sulfur dioxide occurred following direct exposure from leaking application equipment (see Table 28.6). Two California applicator cases suggestive of possible allergic reactions are discussed in the following section. Sensitization Data  Four solid products (10–80% AI) and a liquid product (49.5% AI) caused no sensitization in the Buehler assay. There are no data from standard delayed hypersensitivity assays on sulfur dioxide because of its physical properties. Public domain literature contains limited information regarding sensitization from sulfur and its transformation products. Two case reports implicate elemental sulfur as a human contact allergen. Schneider (1978) reported two cases of contact allergy in patients who used medications containing elemental sulfur to treat superficial fungal dermatoses. Both patients had positive patch test reactions to 5% elemental sulfur in various vehicles. Wilkinson (1975) reported the case of a gardener who developed an eczematous eruption involving the elbow flexures and the right hand. He had a positive patch test reaction to 5% sulfur in petrolatum. Gaul (1960) reported a case of suspected sulfur sensitivity from Indiana related to an acne medication containing sulfur and resorcin. The patient was patch negative to resorcin but had a positive reaction to 2% sulfur in petrolatum. A control series was not reported by Schneider, Wilkinson, or Gaul. Gregorczyk and Swieboda (1968) described 15 cases of desquamative dermatitis among 425 Polish sulfur miners in which irritant dermatitis due to elemental sulfur may have played a part. Reactions to sulfur were also the most common positive patch tests in a pilot study of California nursery workers (O’Malley and Rodriguez, 1998a,b). Positive patch tests to 1% sulfur in butanol occurred in 5 of 39 nursery workers tested compared to 1 of 21 controls. The positive tests were found most often in workers who had directly handled pesticides. However, the workers studied had a limited knowledge of the compounds they handled, and the positive reactions were not clearly relevant to prior exposures to sulfur or prior episodes of contact dermatitis. The possibility of irritant or allergic reactions to transformation products of elemental sulfur is suggested by reports of contact dermatitis associated with bathing in green sulfur springs. Lesions characteristically occurred 24 h after bathing, suggesting a possible delayed contact reaction, but none had positive patch tests to samples of the water (Sun and Sue, 1995). These disparate pieces of information suggest active irritants (e.g., sulfuric acid by oxidation or hydrogen sulfide by reduction) or allergens (e.g., sulfites may be formed from elemental sulfur by oxidation). An older study indicated that finely divided sulfur colloid has antibacterial activity

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

(against Brucella), contradicting the hypothesis that sulfur effects are necessarily mediated by soluble oxidized transformation products (Schuhardt et al., 1952). Frank allergic reactions to an oxidized form of sulfur known as sulfites have been identified in a series of 2894 patients with eczematous dermatitis screened with multiple delayed contact allergens (Vena et al., 1994). Positive patch tests to sodium metabisulfite occurred in 50 patients (1.7%). Positive reactions related to occupational contact occurred in 7 patients (a hairdresser, 2 photographers, a wine producer, an agronomist, a carpenter, and a worker employed in a chemical factory). All 7 had hand dermatitis that flared after contact with sulfites, and it was only possible to identify the source of exposure in 5 of the remaining 43 sulfite-sensitive patients (from topical preparations). Cases of asthma related to ingestion of sulfites also suggest an allergic mechanism. Although pretreatment of patients with sodium cromolyn prior to sulfite challenge prevents the occurrence of bronchospasm, no IgE antibodies to sulfite have been identified. It is therefore possible that sulfites produce release of histamine through nonimmunologic means (Freedman, 1980). California illness cases suggestive of allergy to sulfur 1987-174: A worker complained of rash after mixing, loading, and applying sulfur dust, despite wearing protective equipment. He had a 2-year history of sensitivity to the product. The treating physician believed that the problem was allergic in nature and recommended that he not spray sulfur in the future. 1998-585: An applicator applied sulfur dust to grapevines, despite a history of prior reactions to it. When he woke up the following morning, he had facial swelling and itching that required medical attention that afternoon. O CH3 Cl

O CH C C CH3 N

N

CH3

N Triadimefon

Triadimefon  Similar to other conazole fungicides, triadimefon inhibits the activity of lanosterol 14-demethylase, limiting the biosynthesis of the essential cell wall lipid, ergosterol (Ross et al., 2009). An August 2006 U.S. EPA regulatory action drastically limited the use of triadimefon because of neurotoxicity in animal studies (U.S. EPA, 2006a). California use data for 2007 showed 673 applications, for a total of 872 pounds used on vineyards and nursery crops and for landscape maintenance. This represented a substantial decrease compared to levels used 10 years previously (3096 applications, 12,204 pounds applied in 1997).

763

Physical Properties  Formula C14H16ClN3O2; MW, 293.75; MP, 82°C; log P, 2.77; VP, 1.5  108 mm Hg; solubility in H2O, 260 mg/l at 25°C; other solubilities: moderately soluble in most organic solvents except aliphatics Irritation, Sensitization, and California Illness Data  Five powdered or granular products (1–94.6% AI) and a 1% liquid caused minimal irritation in the Draize assay. The 94.6% solid technical material caused a mild sensitization reaction in the Buehler test (erythema score 1.3 at 24 h and 1.2 at 48 h). There were four cases associated with triadimefon in the handler database, including one case of dermatitis following direct accidental exposure. S

S

S

C N

N TCMTB (thiocyanomethylthiobenzothiazole)

TCMTB (thiocyanomethylthiobenzothiazole)  TCMTB has a nucleophilic cyanate group and an electrophilic substituent, benzothiazole, indicating a probable ability to inhibit with a broad variety of fungal cellular functions. Data for 2007 showed 12,023.52 pounds sold in California for 2007. There was minimal agricultural use reported, suggesting most of the material sold was used as an industrial preservative. Physical Properties  Formula, C9H6N2S3; MW, 238.36; MP, 82°C; log P, 3.30; VP, 3.12  107 mm Hg; solubility in H2O, 125 mg/l at 24°C; other solubilities: soluble in most organic solvents Irritation, Sensitization, and California Illness Data  A mixture of 15.6% TCMTB and 5.3% bis(trichloromethyl) sulfone caused corrosion in the Draize assay. There was no sensitization study available for review from California registration data for TCMTB. The U.S. EPA reregistration eligibility decision (RED) for TCMBTB showed that the 80% AI caused delayed contact hypersensitivity in guinea pigs when induced and challenged by a 40% w/v aqueous concentration of AI. The MBT transformation is a wellrecognized dermal sensitizer. There were no cases associated with its use in the handler database. Cl

O O N

Cl

CH3 O

CH2CH3

Vinclozolin

Vinclozolin  Vinclozolin has a mechanism similar to that of iprodione, affecting cell wall synthesis. Specific enzyme targets include NADPH-cytochrome c reductase (Choi et al.,

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1996). California use data for 2007 showed 109 applications, for a total of 390 pounds used on nursery crops and for landscape maintenance. Physical properties  Formula, C12H9Cl2NO3; MW, 286.11; MP, 108°C; BP, 131°C at 0.05 mm Hg; log P, 3.10; VP, 1.2  107 mm Hg; solubility in H2O, 1000 mg/l; other solubilities (g/kg at 20°C): ethanol 14, acetone 435, ethyl acetate 253, cyclohexane 9, diethyl ether 63, benzene 146, xylene 110, cyclohexanone approximately 540, and chloroform 319 Irritation, Sensitization, and California Illness Data  A 28.3% liquid and a 50% wettable powder caused minimal irritation in the Draize assay. A U.S. EPA memorandum indicated that the 50% wettable powder caused sensitization in the GPMT (U.S. EPA, 1981). There were no cases associated with use of vinclozolin in the handler database.

28.2.4  Fumigants and Biocides 28.2.4.1  Phosphine-Generating Fumigants   and Epoxides

 

AIP

H2O, O2 PH3

PH3 +Al(OH)3 O2 PO4+H2O

Aluminum Phosphide, Magnesium Phosphide, Zinc Phosphide, and Phosphine Aluminum phosphide is a fumigant formulated as solid tablets that release phosphine gas on contact with air and water. It is used for both commodity fumigation and rodent control.

Dermal Sensitization  Animal sensitization and irritation data were not available for review. Dermatitis cases have occurred in handlers following application (84-2184) and contact with partially spent dust.

O Ethylene oxide

Ethylene Oxide­  Ethylene oxide is a reactive electrophile used as a commodity fumigant in food processing and in hospital sterilization equipment. Structurally, it is related to epichlorochydrin and epoxypropane (Birnie and English, 2006). For 2007 CDPR data showed 3,926,035 pounds sold in California. Agricultural use data for 2007 showed minimal use. Physical Properties  Formula, C2H4O; MW, 44.06; MP, 111.7°C; BP, 10.7°C, 760 mm Hg; log P, 0.30; VP, 1314 mm Hg; solubility in H2O, miscible in all proportions with water, alcohol, ethers, and most organic solvents Irritation and sensitization data  No dermal irritation or sensitization study was available for review; however, some ethylene oxide products are labeled as dermal sensitizers. Numerous case reports have described allergic contact dermatitis in hospital workers handling rubber products and other medical supplies sterilized with ethylene oxide (Alomar and Gimenez Camarasa, 1981; Alomar et al., 1981; Fisher, 1988; Hanifin, 1971; Romaguera and Grimalt, 1980; Romaguera and Vilaplana, 1998; Romaguera et al., 1977; Taylor, 1977). The case described in the handler database (1987-2720) involved a chemical burn following accidental direct contact with ethylene oxide gas.

28.2.4.2  Halogenated Fumigants Cl

Physical Properties­  Aluminum phosphide: Formula, AlP; MW, 57.95; MP, 2550°C; log P, NA; VP, negligible; solubility in H2O, decomposes in water; other physical properties: must be protected from moist air because it reacts readily to produce phosphine Zinc phosphide: Solubility in H2O: practically insoluble, decompose slowly; slightly soluble in carbon disulfide, benzene; practically insoluble in alcohols. Magnesium phosphide: Solubility in H2O: reacts with water, but does not dissolve per se. No data available on solubility in other solvents Phosphine: Formula, PH3; MW, 34.00; BP, 87.7°C; log P, NA; VP, 2.93  104 mm Hg; other solubilities: soluble in alcohol, ether, and cuprous chloride solution

Cl

H C C C H H

H

1,3 Dichloropropene

Dichloropropene  Dichloropropene is a fumigant and a biocide that inhibits target cell metabolism at multiple sites. For 2007, California agricultural use data showed 2021 applications, for a total of 9,594,517 pounds used for soil treatments prior to planting grain, vineyards, orchards, row crops, nurseries, berry crops, and uncultivated land. Of currently registered products, 14 of 18 contain chloropicrin (14.8–60% concentrations). Physical properties  Formula, C3H4Cl2; MW, 110.97; MP, 50°C; BP, 108°C; log P, 1.82; VP, 34 mm Hg; solubility

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

in H2O, 2800 mg/l at 20°C; other solubilities: miscible with hydrocarbons, halogenated solvents, esters, and ketones Irritation and Sensitization Data  Dichloropropene mixtures with chloropicrin (see Table 28.6) caused corrosion in the Draize assay. No animal sensitization studies were available for review, but the products registered in California are registered as sensitizers. A case reported by Nater and Gooskens (1976) describes an allergic reaction to a mixture of dichloropropane and dichloropropene (DD mixture), identifying dichloropropene as the most likely allergen. Bousema et al. (1991) and van Joost and de Jong (1988) reported cases of sensitization in a manufacturing operation in The Netherlands. The latter report included a control series of 20 patients negative to a 0.05% test concentration. Another case involved an Italian applicator initially exposed through accidental direct contact who appeared to have a typical case of irritant dermatitis. Following return to work, reexposure, even when wearing complete protective personal equipment, provoked recurrent episodes of dermatitis. Open patch testing with 1% dichloropropene in petrolatum proved negative, but closed patch testing produced a 2 reaction by 48 h. The same concentration produced negative responses in five control subjects (Corrazza et al., 2003). California Illness Data  There were 13 cases associated with dichloropropene in the handler database and 5 additional cases associated with the D-D mixture (see Table 28.6). All 18 cases resulted from direct accidental contact, frequently causing either dermatitis or a frank chemical burn (see sample cases listed in Table 28.6). Case 1988-462 involved a mixer/loader who had drops of dichloropropene fall on his back and noticed a burning sensation at the time. Two weeks later, a rash developed on his back in the area of contact, suggesting possible sensitization. No patch testing was carried out. Br Br Ethylene dibromide

Ethylene Dibromide  Registration of ethylene dibromide (EDB) was cancelled during the 1980s because of concern about its carcinogenicity in rodent bioassays. Prior to the cancellation, it was used as a broad-spectrum fumigant in a manner similar to current use of dichloropropene. Physical Properties  Formula, C2H4Br2; MW, 187.86; MP, 9.97°C; BP, 131–132°C; log P, 1.96; VP, 11.2 mm Hg; solubility in H2O, 4310 mg/l at 30°C; other solubilities: benzene, carbon tetrachloride, and carbon disulfide Irritation and Sensitization Data  No dermal irritation or sensitization data were available for review. However, a

765

fatal case of EDB poisoning (during a confined space entry) reported by Letz et al. (1984) (with blood bromide levels of 380 mg/l; reference level, 4 mg/l) was accompanied by erythema and blisters that had appeared on the trunk and legs 24 h after initial exposure. A co-worker, who also died, had a blood bromide level of 830 mg/l and did not have evidence of burns on the skin when he expired approximately 5 h after onset of exposure (Letz et al., 1984). California Illness Data  There were four cases associated with use of EDB in the handler database, all involving direct accidental exposure (see Table 28.6). H H C

Br

H Methyl bromide

Methyl Bromide  Methyl bromide is a volatile fumigant used as a structural, soil, and commodity fumigant. Its mode of action is not completely understood but probably relates to methylation of enzymes, nucleic acids, and other macromolecules in target organisms (fungi, plant seeds, and nematodes). Despite international controls related to its potential effects on atmospheric ozone, it is still used as a soil fumigant. California use data for 2007 showed 3506 applications, for a total of 6,438,044 pounds used as a commodity fumigant and as a pre-plant treatment for berries (1180 applications, 2,676,240 pounds for strawberries) and other crops. Physical Properties  Formula, CH3Br; MW, 94.94; MP, 93.66°C; BP, 3.5°C; log P, 1.19; VP, 1620 mm Hg; solubility in H2O, 13.4 g/l; other solubilities: very soluble in acetone, benzene, ethyl ether, and ethanol Irritation and Sensitization Data  It is corrosive in the Draize test, but no dermal sensitization data were available for review. California Illness Data  There were 76 cases associated with methyl bromide in the handler database, following episodes of direct exposure to the skin from leaking application equipment. In the 3 sample cases shown in Table 28.6, exposures were aggravated by failure to promptly decontaminate the exposed area and failure to wear chemicalresistant footwear. H H C

l

H Methyl iodide

Methyl Iodide  Methyl iodide is a prospective replacement for methyl bromide. It currently has a U.S. EPA registration but no registration in California. Based on its

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polarizability relative to that of lower molecular weight halogens, it is anticipated to be a stronger alkylating agent than methyl bromide. Physical Properties  Formula, CH3I; MW, 141.94; MP, 66.5°C; BP, 42.5°C; log P, 1.51; VP, 4.05  102 mm; solubility in H2O, 1.39  104 mg/l at 20°C; other solubilities: miscible in alcohol, ether Irritation and Sensitization Data  The 99.7% technical liquid and a mixture of 98% methyl iodide and 2.19% chloropicrin caused severe irritation in the Draize assay. A mixture of 98% methyl iodide and 2% chloropicrin caused corrosion. The 99.7% technical liquid caused sensitization in the GPMT. Two products (98% AI, 2% chloropicrin and 25% AI, 75% chloropicrin) caused sensitization in the Buehler test. Cl Cl C

NO2

Cl Chloropicrin

Chloropicrin  Chloropicrin used in concentrations greater than 2% is considered an active biocidal ingredient rather than a warning agent. Similar to other fumigants, it interferes with target cell chemistry at multiple sites. Physical Properties  Formula, CCl3NO2; MW, 164.38; MP, 64°C; BP, 112°C, 757 mm Hg; log P, 2.09; VP, 24 mm Hg at 25°C; solubility in H2O, 13.4 g/l; other solubilities: very soluble in acetone, benzene, ethyl ether, and ethanol Irritation and Sensitization Data  The 99.5% technical liquid caused corrosion in the Draize assay. There were no dermal sensitization studies available for review. Corrosion associated with liquid chloropicrin was reported during World War I, when it was used as a chemical weapon (Underhill, 1919). California cases involving mixtures of chloropicrin are reported in conjunction with other active ingredients.

Br Propargyl bromide

Propargyl Bromide  Propargyl bromide has been proposed as an alternative to methyl bromide for some fumigant applications (Duniway, 2002; Ruzo, 2006), but it is not currently registered in the United States. Physical Properties  Formula, CH3Br; MW, 94.94; MP, -93.66°C; BP, 3.5°C; log P, 1.19; VP, 1620 mm Hg; solubility in H2O, 13.4 g/l; other solubilities: very soluble in acetone, benzene, ethyl ether, and ethanol Irritation Data  Although no animal irritation or sensitization data were available for review, propargyl bromide

is reported as a skin and mucous membrane irritant from secondary sources (National Library of Medicine, 2009). It also possesses reactive elements identified in both the SICRET (alkynes) and DEREK models (reactive halogens and unsaturated olefin).

28.2.4.3  Methyl Isothiocyanate   (MITC)-Generating Fumigants S CH3NH

C

Na+ . 2 H2O

H2O

CH3N

C S

MITC

S

+ CH3N

Metam-sodium

C O

MIC - up to 4% of MITC concentration

+

CH3 NH2 Methylamine

+ CS2 + H2S Carbon disulfide Hydrogen sulfide

MITC Generators  Methyl isothiocyanate (MITC) is a nucleophile that acts as a nonspecific enzyme inhibitor (Roberts et al., 1998). It is the active biocide generated from the fumigants dazomet, metam sodium, and metam potassium. Other products may include methylamine, carbon disulfide, and hydrogen sulfide. Small amounts of methyl isocyanate may also be formed. MITC dissipation from treated acreage depends on the rate of formation from the parent compound, soil type, postapplication water treatments, wind speed, and temperature. S

S CH3

N

N CH3

Dazomet

Physical Properties  Formula, C2H3NS; MW, 73.11; MP, 36°C; BP, 119°C; log P, 0.94; VP, 3.54 mm Hg; solubility in H2O, 7600 mg/l; other solubilities: soluble in ethanol, methanol, acetone, cyclohexanone, dichloromethane, chloroform, carbon tetrachloride, benzene, xylene, petroleum ether, and mineral oils Dazomet  Dazomet is a fumigant that releases MITC as it breaks down in soil. Other by-products include formaldehyde, monomethylamine, and hydrogen sulfide. In acid soils, carbon disulfide may also be released (U.S. EPA, 2008). For 2007, agricultural use reporting in California showed 60 applications, for a total of 37,537 pounds used principally for nursery crops and landscape maintenance (as a pre-plant treatment). Many products are registered only for use as a biocide in water treatment systems: 2007 data showed 239,092 pounds of dazomet sold in California.

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

Physical Properties  Formula C5H10N2S2; MW, 162.28; MP, 106–107°C; log P, 0.15, pH 7; VP, 2.8  106 mm Hg; solubility in H2O, 1.2 g/l at 25°C; other solubilities (g/100 ml): isopropanol 0.5, xylene 1.1, ethanol 3.0, ethylene glycol 3.0, dioxane 8.0, acetone 13.1, trichloroethylene 30.0, and chloroform 30 Irritation Data  A 24% liquid product caused corrosion in the Draize assay. A 20% liquid product and a 98.5% solid caused minimal irritation. Public domain literature also contains documentation of irritant effects of dazomet. Seven cases of bullous dermatitis following accidental direct contact with an agricultural formulation of dazomet were described from France in 1993 (Garnier et al., 1993). Similar to cases associated with fumigants in California, contamination of clothing, gloves, and footwear caused most of the cases. Surveillance information on pesticide dermatitis cases from Japan showed 21 cases associated with dazomet, reported between 1975 and 2000. However, there were limited clinical details available about the nature of the cases (Horiuchi et al., 2008). Sensitization Data  No animal studies were available for review, but the agricultural formulations of dazomet has been reported as a sensitizer in the public literature (Black, 1973; Richter, 1980). Sensitization associated with its use as an industrial biocide or preservative has also been reported (Emslie, 1993; Warin, 1992). California Illness Data  There were two cases associated with dazomet in the handler database, including a case related to contact with dazomet trapped underneath work clothing (case 1990-2448). Metam Sodium  Metam sodium, a soil fumigant and nematocide, is also effective against weeds and soil fungi. The reaction of metam with water produces MITC, carbon disulfide, hydrogen sulfide, and methylamine. For 2007, California agricultural use data showed 1510 applications, for a total of 9,897,299 pounds used prior to planting row crops (limited acreage), berries, carrots (372 applications, 4,457,632 pounds), and nursery crops. Physical Properties, Metam-sodium  Formula, C2H4NS2. Na; MW, 129.18; MP, decomposes without melting; log P, 1 at 25°C; VP, 4.53  109mm Hg; solubility in H2O, 7.22  105 mg/l at 20°C; stable in concentrated aqueous solution but decomposes in dilute aqueous solution; other solubilities: moderately soluble in alcohol Physical Properties, MITC:  Formula, C2-H3-N-S; MW, 73.11; MP 36°C, BP 119°C; log P, 0.94, 25°C; VP, 4.53 3 1029 mm Hg; solubility in H2O, 7600 mg/l at 20°C;

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Irritation Data  Five liquid products (32.58–43.8% AI) caused corrosion in the Draize assay. Three liquid products (32.7–42.2% AI) were nevertheless reported to cause minimal irritation. Cases of contact dermatitis associated with metam sodium have been reported in several jurisdictions throughout the world. In Germany, the cases stemmed from use of metam in the production of root vegetables (Jung, 1975; Jung and Wolff, 1970a,b; Wolff and Jung, 1970). Cases of dermatitis were also reported from workers wading into the Sacramento River to clean up metam sodium spilled into the river following a train derailment near Dunsmuir, California, in July 1991 (Koo et al., 1995). A case of metam sodium dermatitis in an applicator in the state of Washington has also been reported (O’Malley, 1997). Sensitization Data  A 42% liquid concentrate was tested at 10% dilution for challenge and 1% for dilution. No control animals were reported tested, but there was an increase in dermal response at challenge, indicating probable sensitization. A clinical report from Germany described nine cases of irritant dermatitis related to MITC exposure from either metam sodium or dazomet. Patching test with 0.05% metam sodium in water was positive in eight patients. The authors described the test concentration as nonirritating but did not report use of a control group (Richter, 1980). California Illness Data  There were 59 cases in the California handler database, principally episodes of dermatitis following cases of accidental direct exposure. Sample cases involving chemical burns and secondary cellulitis are listed in Table 28.6.

28.2.5  Herbicides 2 N

+

N

Diquat dibromide

2 BrCH3 +N

. N+ CH3 2 Cl

Paraquat dichloride

28.2.5.1  Bipyridyls Bipyridyls (diquat and paraquat) disrupt photosystem I in photosynthesis (Paraquat Information Center, 2009) but have multiple sites of action in animal as well as plant cells. Diquat  For 2007, California agricultural use data showed 3220 applications, for a total of 70,047 pounds used on grains, nursery crops, potatoes, and rights of way and for landscape maintenance. Physical Properties  Formula, C12H12N2.2Br; MW, 344.05; MP, 337°C; log P, 4.60; VP, 1  107 mm Hg; solubility

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in H2O, 708,000 mg/l at 20°C; other solubilities: slightly soluble in alcohols and hydroxylic solvents; practically insoluble in nonpolar organic solvents

28.2.5.2  Acetamides and Chloracetanilides CH3 N

Irritation Data  A liquid mixture of 2.3% diquat with oxyfluorfen, dicamba, and fluazifop-p-butyl caused severe corrosion in the Draize assay. Three products (2.3–37.3% AI) caused moderate irritation and four (0.23–8.35%) caused minimal irritation.

CH2OCH3 COCH2Cl

CH2CH3

CH2CH3 Alachlor

Acetochlor CH3 CH3

Sensitization Data  Five liquid products (2.3–37.2% AI) did not cause sensitization in the Buehler test.

O N Cl

California Illness Data  There were 22 cases associated with diquat in the handler database, principally involving direct accidental contact with diquat. One of the sample cases shown in Table 28.6 involved failure to decontaminate shoes after they became soaked with diquat and subsequent prolonged contact (case 1986-1498) similar to irritation of lower extremities caused by fumigants. Paraquat  Paraquat is a contact herbicide and dessicant used to control weeds on a variety of grain, vegetable, and fruit crops. Physical Properties  Formula, C12H14N2; MW, 186; BP, at 760 mm Hg, decomposes at 175–180°C; log P, 4.22; VP, approximately 0 mm Hg at 20°C; solubility in H2O, soluble in water; other solubilities: practically insoluble in organic solvents Irritation Data  Two liquid products (22.3 and 36.1% AI) caused moderate irritation in the Draize test. A 43.8% liquid caused minimal irritation. Skin injury associated with application of paraquat and with its misuse has been reported from many areas of the world (Angelo et al., 1986; Botella et al., 1985; Cooper et al., 1994; Gamier et al., 1994; George, 1989; Horiuchi and Ando, 1980; Howard, 1979; Li, 1986; Peachey, 1981; Sugaya, 1976; Swan, 1969; Vilaplana et al., 1993; Villa et al., 1995). Skin injury in most cases has not been associated with systemic effects of paraquat but has occasionally been described. Sensitization Data  Two liquid products (22.3 and 37.1% AI) caused no sensitization in the Buehler test. California Illness Data  There were 35 cases associated with paraquat in the California handler database. The sample cases described in Table 28.6 involved a mild reaction following direct contact with a dilute paraquat spray (1983-480).

CH2CH3 CH2OCH3 N COCH2Cl

O CH3

Butachlor

O Cl

CH2 C N

CH3 CH3 CH CH2OCH3 N COCH2Cl CH2CH3 Metolachlor CH3

CH2CHCH2 CH2CHCH2

Allidochlor

N

CH3 CH C CH2Cl O

Propachlor

Acetanilides are selective preemergence herbicides that interfere with root development by electrophilic inhibition of the fatty acid synthetic enzyme elongase (Hock and Elstner, 2004). The skin effects of these compounds are quite similar, reflecting the high degree of similarity in their chemical structures. Butachlor is a rice herbicide currently used in Asia, but it is not registered in the United States (Ware and Whitacre, 2004). Acetochlor  Acetochlor is not currently registered in California. It is used extensively on crops in the Midwest. For example, use data for 2008 showed 4 million pounds of acetochlor applied in Minnesota (Minnesota Department of Agriculture, 2008). Physical Properties  Formula, C14H20ClNO2; MW, 269.8; MP, 0°C; log P, 3.03; VP, 3.4  108 mm Hg; solubility in H2O, 233 mg/l; other solubilities: soluble in alcohol, acetone, toluene, and carbon tetrachloride Irritation and Sensitization Data  A U.S. EPA summary memorandum indicated that technical acetochlor (95.4%) and an 88% EC caused sensitization with a study protocol that was not specified (U.S. EPA, 1987). Alachlor  U.S. EPA data for 1993–1995 showed approximately 10 million pounds applied in the United States annually on corn, soybeans, sorghum, ornamentals, peanuts, and sunflowers (U.S. EPA, 1998a). For 2007,

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

California use data showed 76 applications of alachlor, for a total 3911 pounds used on beans and corn. Physical Properties  Formula, C14H20ClNO2; MW, 269.768; MP, 40–41°C; BP, 135°C at 0.3 mm Hg; log P, 3.52; VP, 2.20  105 mm Hg; solubility in H2O, 240 mg/l; other solubilities: soluble in diethyl ether, acetone, benzene, chloroform, ethanol, and ethyl acetate Irritation Data  A 42.2% EC caused minimal irritation in the Draize assay. A dermal irritation study on a 45% emulsifiable formulation showed irritation persisting to the conclusion of the study (at 72 h). Sensitization Data  U.S. EPA registration data indicated that alachlor is a sensitizer in the guinea pig test (test protocol not specified) and in the human repeated insult patch test model. Contact sensitization has been reported in the public domain literature by Won et al. (1993). The case involved an applicator who failed to decontaminate her clothing after accidental contact on her lower extremities. Patch tests were positive to 0.1 and 0.2% alachlor but negative to the standard allergen series and to 0.1 and 0.2% paraquat. No reactions were noted in 10 control subjects using 0.05, 0.1, and 0.2% alachlor, except for one who showed a late reaction 20 days after patch testing (a probable instance of de novo sensitization). California Illness Data  There was one case associated with the use of alachlor in the handler database, involving a suspected allergic reaction (1984-537). Allidochlor­  Allidochlor was canceled by the U.S. EPA and by the state of California in 1985. It is historically important as the first acetanilide skin sensitizer and skin irritant. Physical Properties  Formula, C8H12ClNO; MW, 173.64; MP, 145°C; BP, 156°C at 0.5 mm Hg; log P, 4.5; VP, 2.90  106 mm Hg; solubility in H2O, 20 mg/l at 20°C; other solubilities: soluble in alcohol, hexane, and xylene Irritation and Sensitization Data  Allidochlor was reported as moderately irritating to the skin in public domain literature (Morgan, 1982). A 1996 report described three cases of dermatitis related to allidochlor, all involving irritant reactions following accidental direct exposure. In two of the cases, patch testing was carried out, with documented  reactions to allidochlor. The report did not describe testing of control subjects (Spencer, 1966). Butachlor  Butachlor is used as a rice herbicide in Asia. There are no data regarding levels of use available for review.

769

Physical Properties  Formula, C17H26ClNO2; MW, 311.9; MP, 145°C; BP, 92°C at 2 mm Hg; log P, 4.5; VP, 9.40  103 mm Hg; solubility in H2O, 20 mg/l at 20°C; other solubilities: soluble in most organic solvents, including diethyl ether, acetone, benzene, ethanol, ethyl acetate, and hexane Irritation and Sensitization Data  Per public domain report available from the registrant, technical material caused moderate irritation in the Draize assay. Technical butachlor also caused sensitization in the Buehler test (Monsanto, 1999). A case of possible IgE-mediated allergic hepatitis associated with dermal exposure to dermatitis has also been described in a worker from India (Daryani et al., 2007). Metolachlor  California use data for 2007 showed 37 applications, for a total of 2366 pounds used on beans, corn, nursery plants, peas, tomatoes, and rights of way and for landscape maintenance. Physical Properties  Formula, C15H22ClNO2; MW, 283.80; MP, 62.1°C; BP, 100°C at 0.001 mm Hg; log P, 3.13; VP, 3.14  105 mm Hg; solubility in H2O, 530 mg/l at 20°C; other solubilities: soluble in most organic solvents Irritation and Sensitization Data  The Draize test demonstrates that an 85.1% formulation of metolachlor is a moderate dermal irritant. A formulation containing 79% metolachlor also caused sensitization in the Buehler assay. Propachlor Physical Properties  Formula, C11H14ClNO; MW, 211.7; MP, 77°C; BP, 110°C at 0.03 mm Hg; log P, 2.18; VP, 7.4  104 mm Hg; solubility in H2O, 580 mg/l; other solubilities: soluble in common organic solvents except aliphatic hydrocarbons Irritation and Sensitization Data  No California registration data were available for review. Public domain literature indicated that propachlor causes severe irritation of the eye and skin (International Program on Chemical Safety, 1992). However, no details were included regarding circumstances of exposure or the formulations involved. Possible sensitization was described in a report from The Netherlands (Bruynzeel et al., 1993). In a series of 19 dermatitis cases among bulb growers, patch testing showed 5 cases of 1 reaction to 1% aqueous propachlor, considered to be marginally irritant, and 1 case with a 3 reaction, probably related to sensitization.

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28.2.5.3  Nitroaniline Compounds H 7 C 3 N C 3H 7 NO2

NO2

CH3CH2 N CH2CH2CH3 NO2

NO2

CF3

CF3

Trifluralin

Benefin

CH3 CH3CH2

N CH2 C CH2

NO2

NO2

CF3 Ethalfluralin

H 7C 3 N H 7C 3 NO2 NO2

SO2NH2 Oryzalin

Nitroaniline compounds function principally by inhibiting polymerization of tubulin, affecting mitosis of plant cells (Hock and Elstner, 2004). The effect is similar to that of the antimitotic drug colchicine. Other biochemical effects include inhibition of DNA and RNA synthesis and inhibition of reoxidation of reduced plastoquinone during photosynthesis. Metabolic uncoupling has also been described as an in vitro effect at concentrations of 10–100 M (e.g., of trifluralin). Because of their mode of action, they do not affect established weeds. Benefin (benfluralin)  U.S. EPA data from 2004 indicated that 700,000 pounds were used annually in the United States on commercial and residential turf, with lower amounts on alfalfa, lettuce, clover, nonbearing fruit and nut trees, ornamentals, rights of way, fence rows/hedgerows, and conifers U.S. Environmental Protection Agency (EPA) (2004). R.E.D. Facts, Benfluralin. Available online at: http://www. epa.gov/oppsrrd1/REDs/factsheets/benfluralin_fs.pdf. California use data for 2007 showed 365 applications, for a total of 11,604 pounds used on alfalfa, lettuce, and nursery crops and for landscape maintenance. Physical Properties  Formula, C13H16F3N3O4; MW, 335.3; MP, 148–149°C; BP, 65–66.5°C; log P, 5.29; VP, 6.5  105 mm Hg; solubility in H2O, 0.1 mg/l; other solubilities: very soluble 25 g/100 ml at 25°C in acetone, acetonitrile, chloroform, dimethylformamide, dioxane, methyl ethyl ketone, and xylene. Irritation and Sensitization Data  The technical liquid (95.64% AI) caused moderate irritation in the Draize test. A 96.6% solid caused severe irritation. A 50% dry flowable formulation caused sensitization in the Buehler test (U.S. EPA, 1985). California Illness Data  There was one case associated with benefin in the handler database, associated with direct

accidental contact during an application (case 1988-29, Table 28.6). Ethalfluralin  California agricultural use data for 2007 showed 538 applications, for a total of 36,243 pounds used principally on sunflowers, beans, and cucumbers. Physical Properties  Formula, C13H14F3N3O4; MW, 333.3; MP, 55–56°C; BP, decomposes at 256°C; log P, 5.11; VP, 8.8  105 mm Hg; solubility in H2O, 0.3 mg/l; other solubilities: acetone, acetonitrile, benzene, chloroform, dichloromethane, and xylene 500 g/l. Irritation and Sensitization Data  An 18.5% liquid mixture with 5.7% clomazone caused minimal irritation in the Draize assay. The same formulation showed no evidence of sensitization in the Buehler assay, but the study was judged inadequate because of the absence of reaction in positive controls treated with MBT. Summary data from the U.S. EPA indicated that ethalfluralin caused sensitization in the GPMT but did not cause sensitization in the Buehler method. California Illness Data  There was one case associated with ethalfluralin in the handler database – facial dermatitis following accidental direct exposure (case 1990-1832, Table 28.6). Oryzalin  California agricultural use data for 2007 showed 10,974 applications, for a total of 656,439 pounds used on orchards, berry crops, vineyards, nurseries, and rights of way. Physical Properties  Formula, C12H18N4O6S; MW, 346.36; MP, 141°C; BP, decomposes at 265°C; log P, 3.73; VP, 9.75  109 mm Hg; solubility in H2O, 2.5 mg/l; other solubilities: soluble in ethanol; practically insoluble in hexane. Irritation and Sensitization Data  Four products (96.55% powder, 41% liquid, 40.4% liquid, and a mixture of 1% oryzalin and 1% benefin) caused minimal irritation in the Draize assay. The 40.4% liquid product caused no sensitization in the Buehler assay. California Illness Data  Two cases associated with oryzalin were included in the pesticide handler database, both occurring following accidental direct contact (1984-51 and 1984-272, Table 28.6). Pendimethalin  California agricultural use data for 2007 showed 15,098 applications, for a total of 1,124,396 pounds used on alfalfa, almonds, pistachios and other orchard crops, corn, vineyards, nurseries, and row crops and for landscape maintenance. California sales data showed 1,653,146 pounds sold in 2007, considerably exceeding the reported agricultural use. Nonagricultural use derives from mixed formulations in “weed and feed” lawn care products. Physical Properties  Formula, C13H19N3O4; MW 281.31; MP, 281.31°C; BP, 330°C; log P, 5.18; VP, 3  105 mm Hg; solubility in H2O, 0.3 mg/l; other solubilities: readily soluble in benzene, toluene, chloroform, and dichloromethane.

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

Irritation and Sensitization Data  Two products (a 38.7% suspension and a 37.4% liquid) caused minimal irritation in the Draize test. The 38.7% suspension caused no sensitization in the Buehler assay. California Illness Data  There were two cases associated with pendimethalin in the handler database, but none following direct accidental contact. Trifluralin  At the time of reregistration in 1996, U.S. EPA data showed 25 million pounds used annually on agricultural crops, with soybeans accounting for 64% and cotton for 19%. Other treated crops included green beans, broccoli, tomatoes, cabbage, sunflowers, dry beans, cauliflower, okra, soybeans, carrots, flax, Brussels sprouts, asparagus, and sweet peppers (U.S. EPA, 1996). Physical Properties  Formula, C13H16F3N3O4; MW, 335.28; MP, 46–47°C; BP, 139–140°C; log P, 5.34; VP, 4.58   105 mm Hg; solubility in H2O, 18.4 mg/l; other solubilities: 100 g/100 ml acetone and 81 g/100 ml xylene Irritation Data  A 50.8% EC caused moderate irritation in the Draize assay. Seven products (0.74% granules–80% dry flowable formulation) caused minimal irritation. Sensitization Data  A product with 2% trifluralin, 0.25% isoxaben, and 0.25% oxyfluorfen caused sensitization in the GPMT. A 43% liquid caused no sensitization in the Buehler test. California Illness Data  There were 14 cases associated with trifluralin in the handler database, including typical cases of dermatitis following accidental direct contact (1993-340 and 1994-565, Table 28.6).

28.2.5.4  Organophosphate-like Compounds O HO

O

C CH2NHCH2 P OH OH

Glyphosate

S SO2 NH CH2 CH2

S

P

O CH(CH3)2 O CH(CH ) 3 2

Bensulide O HO

O _ CH3 C CH2NHCH2 P O + S CH3 OH Sulfosate

C 4H 9 S C 4H 9 S P C 4H 9 S Folex Merphos

CH3

C4 H9 S C 4H 9 S P O C 4H 9 S DEF Tribufos

771

Although most important organophosphates are insecticides, the group includes several herbicidal compounds that do not inhibit cholinesterase to any significant degree and two phosphorothioate compounds (buffos and merphos), used as cotton defoliants, that are weak cholinesterase inhibitors, approximately comparable to malathion (Hayes, 1982). Bensulide  Bensulide acts by inhibiting cell division in root tips (Ware and Whitacre, 2004). It also has some activity as a cholinesterase inhibitor (EXTOXNET, 2009). California use data for 2007 showed 7050 applications, for a total of 258,164 pounds used on row crops, nursery crops and for landscape maintenance. Physical Properties  Formula, C14H24NO4PS3; MW 397.54 MP, 34.4°C; BP, 397.52°C; log P, 4.20; VP, 8.0   107 mm Hg; solubility in H2O, 25 mg/l; other solubilities: miscible with acetone, ethanol, methyl isobutyl ketone, and xylene Irritation, Sensitization, and California Illness Data  Six bensulide products (solid and liquid formulations, 3.6–92.5% AI) caused minimal irritation in the Draize assay. Neither a 93.8% liquid nor a 12.5% granular product caused sensitization in the Buehler test. There were two cases associated with bensulide in the pesticide handler database, including one case of dermatitis following accidental direct exposure (1983-1770, Table 28.6). Glyphosate and Sulfosate  Glyphosate is a nonselective herbicide with extensive agricultural and nonagricultural uses. It causes inhibition of synthesis of the amino acids phenylalanine and tyrosine; because these amino acids are obtained from dietary sources in mammals, glyphosate consequently has low human systemic toxicity (Ware and Whitacre, 2004). Formulated products containing glyphosate have previously been shown to cause both sensitization (related to an isothiazolin preservative) and irritation (probably related to surfactant content). In addition to possible variations in surfactants and other inert compounds, products may contain one of several variants of the active ingredient. These include the glyphosate trimesium salt, sulfosate, as well as its isopropylamine, sesquisodium, ammonium and diammonium, potassium, and dimethylamine salts. Sales of glyphosate and related products in California for 2007 totaled 14,270,934 pounds. Reported agricultural use for glyphosate and related salts for 2007 included 7,236,787 pounds, for a total of 139,568 applications on orchards, grains, vineyards, rights of way, nursery crops, forage, row crops, and berries and for landscape maintenance.

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Glyphosate Physical Properties  Formula, C3H8NO5P; MW, 169.1; BP, decomposes above 200°C; log P, -3.40; VP, 9.8  108 mm Hg; solubility in H2O, 10.5 g/l at pH 1.9 and 20°C; other solubilities: insoluble in common organic solvents such as acetone, ethanol, and xylene Irritation Data  Most glyphosate-containing products caused minimal irritation in the Draize assay. These included 9 liquid products and 1 solid product (1–62% AI). Four liquid products (28.3–48.8% AI) caused moderate irritation and 1 liquid product containing 48.7% liquid caused severe irritation. Sensitization Data  Four liquid products (41–48.8% AI) caused no sensitization in the Buehler assay. However, a 40.1% liquid product caused sensitization in the LLNA. Test materials diluted 5, 25, and 75% elicited an SI of 1.7-, 9.3-, and 18.5-fold compared to controls. The estimated EC3 for the test product was 7 to 8%. A photoallergic reaction to an isothiazolin preservative present in a glyphosate formulation was reported by Hindson and Diffey (1984a,b). No reaction was noted to the AI. California and Other Illness Data  Glyphosate formulations are probably infrequent sources of dermatitis, given their volume of use, and have been determined experimentally to be no more irritating than detergents contained in topical shampoos (Maibach, 1986). Nevertheless, glyphosate formulations accounted for 204 (53%) of all 384 possible, probable, and definite cases associated with herbicides in the handler database. Merphos (DEF) and Tribufos  Tribufos (Folex) is a phosphotrithioate used as a defoliant to minimize boll rot in cotton and to prepare the plants for mechanical harvesting. The oxidation product omerphos, DEF, is also used as a cotton defoliant. Physical Properties Tribufos: Formula, C12H27OPS3; MW, 314.54; MP, -25°C; BP, 210°C at 750 mm Hg; log P, 5.7; VP, 5.3  106 mm Hg; solubility in H2O, 2.3 mg/l; other solubilities: soluble in aliphatic, aromatic, and chlorinated hydrocarbons and alcohols; completely miscible in dichloromethane, n-hexane, 2-propanol, and toluene Merphos: Formula, C12H27PS3; MW, 298.5; MP, 115– 134°C; BP, 115–134°C (at 0.08 mm Hg); log P, 7.670; VP, 0.08 mm Hg; solubility in H2O, limited; other solubilities: soluble in acetone, ethyl alcohol, benzene, hexane, kerosene, diesel oil, heavy aromatic naphthas, xylene, and methylated naphthalene

Cl

O

Cl

O CH2 Cl

C OH Cl

2,4,5-T

O O CH2

Cl

C OH

2,4-D COOH Cl

Cl

OCH3

OCH2COOH CH3

Cl Dicamba

MCPA CL CH3 Cl

OCHCOOH MCPP

Tribufos Irritation, Sensitization, and California illness data  The 71% liquid caused corrosion in the Draize assay and the 99.7% liquid caused moderate irritation. The sensitization study reviewed did not contain sufficient information to determine whether or not tribufos is an allergen in the Buehler assay. No tribufos cases were listed in the handler database.

28.2.5.5  Phenoxy Herbicides 2,4-D and 2,4,5-T are the prototype phenoxy compounds, controlling broadleaf weeds through effects on plant hormones called auxins. 2,4,5-T was removed from the market in the late 1970s because of contamination with 2,3,7,8-TCDD. The contaminant resulted from hydroxylation of tetrachlorobenzene to produce the trichlorophenol component of 2,4,5-T. 2,4-D is manufactured by chlorination of phenol and does not contain the same dioxin contaminants. It is still used in a broad variety of agricultural and nonagricultural products. CDPR registration data contain compound numbers for 107 compounds related to 2,4-D, including multiple ester derivatives and alkyl amine salts. Formulations often contain mixtures with related compounds such as dicamba, MCPA, and MCPP. Reviewing the irritation and sensitization data for all of these related compounds is beyond the scope of this chapter. Representative data are reviewed here. California use data for 2007 showed 442,107.0 pounds applied, for a total of 19,451 applications on orchards, grain, forage, nurseries, forest and rangeland and for landscape maintenance for 2,4-D, 5 ester derivatives and 6 amine salts. Approximately 1,215,123 pounds of 2,4-D and its derivatives were sold in California during the same year.

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

Physical properties  2,4-D: Formula, C8H6Cl2O3; MW, 221.04; MP 138°C; BP, 160°C; log P, 2.81; VP, 8.25  108 mm Hg at 20°C; solubility in H2O, 677 mg/l at 25°C; other solubilities: 67.3 g/400 ml acetone at 25°C Dicamba: Formula, C8H6Cl2O3; MW, 221.04; MP, 114–116°C; log P, 2.21; VP, 3.4  105 mm Hg; solubility in H2O, 6500 mg/l; other solubilities: solubility in xylene 78 g/l MCPA: Formula, C9H9ClO3; MW, 200.6; MP, 118–119°C; log P, 3.25; VP, 5.90  10-6 mm Hg; solubility in H2O, 630 mg/l; other solubilities (g/100 ml): ether 77, ethanol 153, n-heptane 0.5, toluene 6.2, and xylene 4.9 Irritation Data  Six liquid 2,4-D products (0.2–38.3% AI) and two products containing 2,4-D powder (96.7 and 86% AI) caused minimal irritation in the Draize assay. A liquid product containing 51.9% MCPP, granules containing low concentrations of MCPA (0.82%), MCPP (0.33%), and dicamba (0.06%), and an 86.8% solid formulation of dicamba also caused minimal irritation. 2,4-D products causing moderate irritation included 1) an EC mixture of 5.38% dicamba, 32.5% 2,4-D-2-ethylhexyl ester, and 15.9% 2,4-DP-P, isooctyl ester, 2) a liquid containing 11.6% DMA salt of 2,4-D and 11.6% DMA salt of MCPP, and 3) granules containing 1.37% 2,4-D, 1.37% MCPP, and 0.55% dicamba. A soluble powder with 20% MCPP, 50% MCPA, and 5% dicamba caused severe irritation. Sensitization Data­  A liquid product containing 19.6% 2,4-D and a solid formulation with 86.8% dicamba both caused sensitization in the Buehler assay. Sensitization has also been reported in the public domain literature. Among 30 Indian farmers with chronic contact dermatitis (Sharma and Kaur, 1990) of 2.5 years’ average duration, 3 proved to have contact sensitivity to 2,4-D (1% in petrolatum). Five cases of irritant bullous dermatitis also affected German forestry workers applying a mixture of 2,4,5-T and 2,4-D. Two cases had positive patch test reactions to 0.4% concentration of the product mixture in diesel oil. The first demonstrated 3 reaction at 48 and 72 h. A weakly positive reaction was observed in the second case (Jung and Wolf, 1975). Four 2,4-D products were negative in the Buehler assay. These included 2,4-D powders containing 100 and 86% AI and two granular products with mixed phenoxy acid compounds [1.22% 2,4-D,1.22% MCPP, and 0.081% dicamba; 0.64% 2,4-D, 0.15% R()MCPP, 0.06% dicamba, and 0.19% dithiopyr]. An aqueous concentrate containing DMA salts of MCPA (40.42%), MCPP (15.97%), and dicamba (3.97%) and a similar concentrate containing DMA salts of MCPP (10%) and MCPA (14%) caused no sensitization in the Buehler assay.

773

Cl Cl

O O CH2

Cl

C OH Cl

2,4,5-T

O CH2

Cl COOH Cl

C OH

2,4-D

OCH3 Cl

OCH2COOH

Cl

Dicamba

CH3

CL CH3 Cl

O

MCPA

OCHCOOH MCPP

California Illness Data  There were eight cases associated with phenoxy herbicides in the pesticide handler database, including one case of dermatitis following accidental direct contact with MCPA (see Table 28.6).

28.2.5.6  Pyridine Carboxylic Acids NH2 Cl

Cl

Cl COOH N Picloram

Pyridine carboxylic acids have a planar structure akin to the phenoxy herbicides and a similar effect on hormones (auxins) in broadleaf weeds (Ware and Whitacre, 2004). Examples discussed here are picloram and triclopyr. Picloram­  Picloram is a synthetic auxin (Fuersta et al., 1996) used for control of annual and perennial broadleaf weeds, woody plants, and vines. Agricultural use data for 2007 showed minimal use in California. Physical Properties­  Formula, C6H3Cl3N2O2; MW, 228.46; MP, 218.5°C; log P, 0.30; VP, 7.21  10–11 mm Hg; solubility in H2O, 430 mg/l; other solubilities: organic solvents (g/100 ml at 25°C): acetone 1.98, acetonitrile 0.16, benzene 0.02, carbon disulfide 0.005, diethyl ether 0.12, ethanol 1.05, isopropanol 0.55, kerosene 0.001, and methylene chloride 0.06 Irritation, Sensitization, and Illness Data­  The 20.4 and 38.8% liquid products caused minimal irritation in the Draize assay. No sensitization studies were available for review.

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There was one case associated with use of picloram in the handler database that did not involve direct contact (see Table 28.6). Cl

Cl

Irritation, Sensitization, and Illness Data­  10.3 and 15.2% granular products caused minimal irritation in the Draize assay. The 15.2% granular product was a nonsensitizer in the Buehler assay.

O Cl

OCH2 N Triclopyr

Physical Properties  Formula, C9H17NOS; MW, 187.31; BP, 136.5°C, 10 mm Hg; log P, 3.21; VP, 5.6  103 mm Hg; solubility in H2O, 970 mg/l; other solubilities: miscible with acetone, ethanol, kerosene, 4-methylpentan-2-one, and xylene

C OH

Triclopyr  Triclopyr is a systemic herbicide for control of woody plants, broadleaf weeds, forests, turf, and industrial sites. Physical Properties  Formula, C7H4Cl3NO3; MW, 256.5; MP, 150.5°C; log P, 2.530; VP, 1.26  106 mm Hg; solubility in H2O, 440 mg/l; other solubilities (g/l): acetone 581, acetonitrile 92.1, hexane 0.09, toluene 19.2, dichloromethane 24.9, methanol 665, and ethyl acetate 271 Irritation Data  Five products containing liquid triclopyr or liquid triclopyr butoxyethyl ester (13.6–85.3% AI) caused moderate irritation in the Draize assay. Nine products (eight liquid and one granular formulation, 6.6–97.1% AI) caused minimal irritation. Sensitization Data  Products containing 85.3 and 61.2% liquid triclopyr butoxyethyl ester caused sensitization in the LLNA. An aqueous concentrate with 44.6% triclopyr caused sensitization in the Buehler assay. Products containing 13.6% triclopyr ester, 14.8% triclopyr granules, and a mixture of 33% triclopyr and 12.3% clopyralid caused no sensitization in the Buehler test. A case of dermatitis following direct contact with triclopyr was reported in the pesticide handler database (1988-2832, Table 28.6).

The single case reported in the handler database indicated that occlusion of molinate against the skin inside protective clothing may cause dermal irritation (see Table 28.6). Thiobencarb  California use data for 2007 showed 805 applications, for a total 289,046 pounds applied on rice, with minor use on nursery plants and for landscape maintenance. Physical Properties  Formula, C12H16ClNOS; MW, 257.8; MP, 3.3°C; BP, 126-129°C (at 0.008 mm Hg); log P, 3.4; VP, 2.20E-05 mm Hg; solubility in H2O, 28 mg/l; other solubilities: readily soluble in acetone, ethanol, xylene, methanol, benzene, n-hexane, and acetonitrile Irritation, Sensitization, and Illness Data  All products tested, including 15% granules, 84% EC, and 97.4% liquid, caused minimal irritation in the Draize assay. A granular product with 15% thiobencarb caused no sensitization in the Buehler assay. There were no cases associated with thiobencarb in the pesticide handler database.

28.2.5.8  Triazines

N H7C3NH

CH2

Thiobencarb

CH3S

NHC2H5

N

Cl

C2H5S C N Molinate

N H7C3NH

Cl

N

N

NHC3H7

N

NHC3H7

N Prometryn

H5C2NH

N N

NHC2H5

Simazine

Prometon

Carbamate herbicides do no inhibit cholinesterase but are suspected to inhibit lipid synthesis in plants at the step catalyzed by the enzyme acetyl-CoA elongase (Ware and Whitacre, 2004). They function as preemergent and early postemergent herbicides for control of grasses and broadleaf weeds on rice fields.

N

OCH3

O

O N C S

C2H5

N

N

Atrazine

28.2.5.7  Carbamates C2H5

NHC3H7

Cl

Cl N CH3NH

N N

CH3 NH C CN CH3

Cyanazine

Molinate  California use data for 2007 showed 214 applications, for a total of 75,241 pounds, almost exclusively on rice, with minor uses on nursery crops.

Triazines inhibit photosynthetic electron transport and function as postemergent herbicides. Tolerant plants

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

metabolize the parent compound, but susceptible plants do not (Whitacre and Ware, 2004). Atrazine  Prior to U.S. EPA regulations promulgated in 2003 to limit groundwater contamination, 80 million pounds were used in the United States annually (Donaldson et al., 2002; Ware and Whitacre, 2004). California use data for 2007 showed 378 applications, for a total of 27,546 pounds used on corn, forest timberland, and grains. Physical Properties  Formula, C8H14ClN5; MW, 173; MP, 173°C; log P, 2.61; VP, 2.89E-07 mm Hg; solubility in H2O, 34.7 mg/l; other solubilities (g/kg): DMSO 183, chloroform 52, ethyl acetate 28, methanol 18, diethyl ether 12, and pentane 0.36 Irritation, Sensitization, and Illness Data  Atrazine products containing 1.16% granules, 40.8% liquid, and 33.1% AI mixed with 26.1% metolachlor caused minimal irritation in the Draize assay. The 40.8% liquid caused no sensitization in the Buehler test. There were no cases associated with atrazine in the handler database. Cyanazine  Cyanazine is a selective herbicide used principally on cotton in California and on corn in the grain-producing areas of the United States. Physical Properties  Formula, C9H13ClN6; MW, 240.70; MP, 168°C; log P, 2.22; VP, 1.38E-07 mm Hg; solubility in H2O, 170 mg/l; other solubilities (g/l at 25°C): benzene 15, chloroform 210, ethanol 45, and hexane 15 Irritation, Sensitization, and Illness Data  The 97.3% technical solid caused minimal irritation in the Draize test. An EC mixture containing 22% metolachlor and 22% cyanazine caused no sensitization in the Buehler assay. There were no cases associated with cyanazine in the hand­ ler database. Prometon  Current products are registered for home and garden use. California pesticide use reports for 2007 indicate minimal agricultural use. Physical Properties  Formula, C10H19N5O; MW, 225.3; MP, 91.5°C; log P, 2.99; VP, 2.30E-06 mm Hg; solubility in H2O, 750 mg/l; other solubilities: organic solvents (g/100 ml at 20°C): n-hexane 1.2, methanol 60, cylohexane 4.9, and n-octanol 26 Irritation, Sensitization, and Illness Data  A 1.86% liquid product caused severe irritation in the Draize assay. Two liquid products (2 and 12.5% prometon) and two mixed RTU products (a product containing 3.59% prometon and 1% 2,4-D and a product with an inactive registration

775

containing 2.5% prometon and 1% pentachlorophenol) caused moderate irritation. Three products (a 3.75% RTU liquid, a 45.3% EC, and a 97.3% technical solid) caused minimal irritation. Four products (including a 97% technical solid, a 45.3% liquid, and separate mixtures of prometon with 2,4-D and diquat dibromide) caused no sensitization in the Buehler assay. There was one case associated with use of prometon in the handler database. Prometryn  Prometryn is a selective herbicide with a spectrum of use similar to other triazines. California use data for 2007 showed 2308 applications, for a total of 69,526 pounds used principally on celery and cotton. Physical Properties  Formula, C10H19N5S; MW, 228.57; MP, 119°C; log P, 3.51; VP, 2.00E-06 mm Hg; solubility in H2O, 33 mg/l; other solubilities (g/l): acetone 330, ethanol 140, hexane 6.3, toluene 200, and n-octanol 110 Irritation, Sensitization, and Illness Data  A 97.3% technical solid caused minimal irritation in the Draize assay. A 45% formulation caused minimal reaction in the Buehler test. There were no prometryn-associated cases in the handler database. Simazine  California use data for 2007 showed 12,515 applications, for a total of 538,627 pounds used on orchards, berries, corn, forest land, vineyards, nurseries, rangeland, and rights of way and for landscape maintenance. Physical properties  Formula, C7H12ClN5; MW, 201.66; MP, 226°C; log P, 2.18; VP, 2.21E-08 mm Hg; solubility in H2O, 6.2 mg/l; other solubilities (mg/l at 25°C): ethanol 570, acetone 1500, toluene 130, n-octanol 390, and nhexane 3.1 Irritation, Sensitization, and Illness Data  A 90% dry flowable product, a 28.9% liquid, and a 0.6% RTU formulation caused minimal irritation in the Draize test. A 6.3% formulation caused no sensitization in the maximization test. The single case reported to the handler database involved contamination of the hands with simazine while mixing the material and a secondary dermatitis of the genitalia (case 1983-2394).

28.2.5.9  Urea Herbicides The urea herbicides function by inhibiting the acetolactate synthase step in branched-chain amino acid synthesis (Subramanian et al., 1991; Ware and Whitacre, 2004). They have low systemic toxicity and minimal effects on the skin. For the urea herbicides shown in Table 28.6, only three cases were reported in the pesticide handler database.

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CH3

O Br

N

C CH 3 O

N

CH3

O

Cl

NH C N

CH3

H

N

Diuron

Cl

Bromacil

N

CH3

O NH C NH

S

Thidiazuron COOH O

Cl

O

N

OCH3

S NH C NH N N

OCH3

O

N

S NH C NH

N N

O Cl

OCH3

Halosulfuron

CH3

O

N

O

CH3

Chlorsulfuron N

O

O

N

OCH3

S NH C NH N

O O S O

OCH3 Rimsulfuron

CH3

Bromacil  California use data for 2007 showed 800 applications, for a total of 85,096 pounds used on citrus orchards, and rights of way and for landscape maintenance. Physical Properties  Formula, C9H13BrN2O2; MW, 261.12; MP, 158°C; log P, 2.11; VP, 3.07  107 mm Hg; solubility in H2O, 815 mg/l; other solubilities (g/l at 25°C): ethanol 134, aceto3.07X10-7ne 167, acetonitrile 71, xylene 32, and 3% aqueous sodium hydroxide 88 Irritation, Sensitization, and Illness Data  Three dry flowable products (53% bromacil, 27% diuron; 40% bromacil, 40% diuron; and 80% bromacil) and a dilute mixture of 1.5% bromacil with sodium chlorate and sodium metaborate caused minimal irritation in the Draize assay. The 40% bromacil, 40% diuron product was a nonsensitizer in the Buehler assay. There were no cases associated with bromacil in the handler database. Chlorsulfuron  California use data for 2007 showed 173 applications, for a total of 3,668 pounds used on grains and silage and rights of way and for landscape maintenance.

Physical Properties Formula, C12H12ClN5O4S; MW, 357.78; MP, 176°C; log P, 2; VP, 2.30  1011 mm Hg; solubility in H2O, 2.80E  04 mg/l; other solubilities (at 22°C): 57 g/l acetone, 102 g/l dichloromethane, 10 mg/l hexane, 14 g/l methanol, and 3 g/l toluene Irritation, Sensitization, and Illness Data  The 75% dry flowable caused minimal irritation in the Draize test. There were no cases associated with its use in the handler database. Diuron  California use data for 2007 showed 13,240 applications, for a total of 859,909 pounds used on grains, orchards, row crops, corn, vineyards, cotton, nurseries, rights of way, and uncultivated agricultural land. Physical Properties  Formula, C9H10Cl2N2O; MW, 233.10; MP, 158°C; log P, 2.68; VP, 6.9  108 mm Hg; solubility in H2O, 42 mg/l; other solubilities: very low in hydrocarbon solvents Irritation, Sensitization, and Illness data  Two diuron products caused severe irritation or corrosion in the Draize assay. Both contained the isothiazilone compound octhilinone (2.7% in the product causing severe irritation and 6% in the product causing corrosion). An 81% diuron dry flowable product caused moderate irritation. Three solid products (20–80% diuron) and two liquids with 40% AI caused minimal irritation in the Draize assay. A product containing 7.5% carbendazim, 20% diuron, and 2.7% octhilinone caused sensitization in the Buehler assay. A 40% liquid and a 80% wettable granule product were negative for sensitization. There were no cases associated with diuron in the handler database. Halosulfuron-Methyl  Halosulfuron-methyl is registered for use on a broad variety of grains, row crops, and orchards. It has also been used for weed control on turf and ornamental plants. Total use reported in California agriculture for 2007 was 2818 pounds in 1380 separate applications. Physical Properties  Formula, C13H15ClN6O7S; MW, 434.81; MP, 176°C; log P, 0.02; VP, 1  107 mm Hg; solubility in H2O, 15 mg/l Irritation, Sensitization, and Illness Data  Three powdered products (51.1–98.5% AI) and a granular product containing 12.5% halosulfuron and 55% dicamba caused minimal irritation in the Draize assay. The dicamba, halosulfuron granule is also a nonsensitizer in the Buehler test. There were no cases associated with its use in the handler database. Rimsulfuron  Rimsulfuron is registered for use on a broad variety of grains, row crops, and orchards. Total use reported in California agriculture for 2007 was 2225 pounds in 2255 separate applications.

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Physical Properties  Formula, C14H17N5O7S2; MW, 431.43; MP, 177°C; log P, 0.29; VP,1.13  108 mm Hg; solubility in H2O, 10 mg/l

weeds than grasses. Examples include fluazifop-butyl and flumetsulam (Ferreira et al., 1995; Ware and Whitacre, 2004).

Irritation, Sensitization, and Illness Data  Three solid rimsulfuron products (25% powder, 25% soluble granules, and 98% technical solid) caused minimal irritation in the Draize assay. The 25% granular product was negative for sensitization in the LLNA. There were no cases associated with its use in the handler database.

Fluazifop-butyl  For 2007, California use reports showed 10,192.046 pounds of fluazifop-butyl used in 548 applications to orchard, vineyard, nursery, and roadside rights of way.

Sulfometuron  California data for 2007 indicated that 80% of the 10,021 pounds used was applied to roadside rights of way, with smaller amounts used in greenhouses and for landscape maintenance. Physical Properties  Formula, C14H14N4O5S; MW, 364.40; MP 193-194°C; log P, 0.870; VP, 1.07E-12 mm Hg; solubility in H2O, 70 mg/l Irritation, Sensitization, and Illness Data  A dry flowable product and a granular product, both containing 75% sulfometuron, caused minimal irritation in the Draize assay. The granular product was also negative for sensitization in the Buehler test. Thidiazuron  Thidiazuron is a plant growth regulator and defoliant used on cotton. Physical properties  Formula, C14H14N4O5S; MW, 220.25; MP, 213°C; log P, 0.870; VP, 1.07E-12 mm Hg Irritation, Sensitization, and Illness Data  A mixture of 12% thidiazuron and 6% diuron caused minimal irritation in the Draize test. The same mixture caused no sensitization in the Buehler assay. There were no cases associated with its use in the handler database.

28.2.5.10  Imidazolinones CH3

Irritation, Sensitization, and Illness Data  A 24.9% liquid product caused moderate irritation in the Draize assay; a dilute RTU mixture with 2,4-D and diquat caused minimal irritation. There were 14 cases associated with fluazifop in the pesticide handler database, including 3 resulting from accidental direct contact (see Table 28.6). Flumetsulam  Flumetsulam (triazolopyrimidine sulfonanilide) is a herbicide used only for corn and soybeans. There was no reported use in California in 2007. Physical Properties  Formula, C19H20F3NO4; MW, 325.29; MP, 13°C; log P, 4.5; VP, 4.12E-07 mm Hg; solubility in H2O, 1 mg/l Irritation, Sensitization, and Illness Data  The only irritation and sensitization studies involved a mixture of metolachlor (79.9%) and flumetsulam (2.6%). The mixture was a category II irritant in the Draize test and a sensitizer in the Buehler assay. There were no cases associated with the use of flumetsulam in the handler database.

28.2.5.11  Herbicides of Miscellaneous Structure

O C CHCOO(CH2)3CH3

O

CF3

Physical Properties  Formula, C19H20F3NO4; MW, 383.4; MP, 13°C; BP, 165°C (at 0.02 mm Hg); log P, 4.5; VP, 4.12E-07 mm Hg; solubility in H2O, 1 mg/l; other solubilities: miscible with acetone, cyclohexanone, hexane, methanol, dichloromethane, and xylene

CN

N Fluazifop-butyl

Br

Br

N N N

OH

N

F

Bromoxynil

SO2 NH

Flumetsulam

F

The imidazolinones inhibit amino acid branched-chain biosynthesis; they have more effect against broadleaf

Bromoxynil  Bromoxynil functions by inhibiting electron transport during photosynthesis (Takano et al., 2008). In 2007, a total of 67,433 pounds (octanoate and haptanoate salts) used on barley, cotton, silage, vineyards, onions, and wheat and for landscape maintenance.

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Physical Properties  Formula, C7H3Br2NO; MW, 276.9; MP, 194–195°C; log P, 2.8; VP, 4.72  108 mm Hg; solubility in H2O, 130 mg/l; other solubilities: in water 130 ppm (0.013%) w/v at 20–25°C; in ethanol 7% w/v at 20–25°C; in light petroleum and xylene 1–2% w/v at 20–25°C Irritation, Sensitization, and Illness Data  EC products containing 55 and 33.4% bromoxynil caused minimal irritation in the Draize assay. However, the 33.4% EC product caused sensitization in the Buehler assay. There were two cases associated with bromoxynil in the handler databases, both following accidental direct contact. N C Cl

Cl

Dichlobenil

Dichlobenil  Dichlobenil is used for selective weed control in growing cranberries, ornamental flowers, orchard fruit, vineyards, and turf. It is available in wettable powder and granular formulations. Physical properties  Formula, C7H3Cl2N; MW, 172.01; MP, 144.5°C; BP, 270°C; log P, 2.74; VP, 0.00101 mm Hg; solubility in H2O, 14.6 mg/l; other solubilities: solubility in solvents (w/v) (approximate): acetone 5% at 8°C, benzene 5% at 8°C, cyclohexanone 7% at 15–20°C, ethanol 5% at 8°C, furfural 7% at 8°C, methylene chloride 10% at 20°C, methylethylketone 7% at 15–20°C, tetrahydrofuran 9% at 8°C, toluene 4% at 20°C, and xylene 5% at 8°C Irritation, Sensitization, and Illness Data­  The 98.8% technical solid, a 15.12% liquid, and a 0.55% dust caused minimal irritation in the Draize test. A 15.3% aqueous concentrate was a sensitizer in the GPMT. A case of dermatitis following exposure to a mixture of dichlobenil and dichlorobenzoyl chloride used in chemical synthesis was reported (de Boer and van Joost, 1988). Dichlobenil was discussed as a possible chemical irritant, but sensitization was associated only with dichlorobenzoyl. Cases of chloracne associated with the manufacture of dichlobenil have also been described. The specific contaminant involved was not identified (Deeken, 1974). Limited use of dichlobenil is reported in California, and there were no cases associated with its use included in the handler database. CH3

CH3 CH CH2

CH3S

CO SCH 3

CO CF3

N Dithiopyr

CF2H

Dithiopyr  Dithiopyr functions as preemergent herbicides; its biochemical mechanism is inhibition of microtubule assembly (Ware and Whitacre, 2004). California use data for 2007 showed 29 applications, for a total of 10,026 pounds applied for landscape maintenance and to nurseries and rights of ways. Physical properties Formula, C15H16F5NO2S2; MW, 401.409; MP, 65°C; log P, 4.75; VP, 4.00E-06 mm Hg; solubility in H2O, 1.4 mg/l Irritation, Sensitization, and Illness Data  The 22.9% EC caused severe irritation in the Draize assay; the 13.5% liquid, 28.4% solid, and 91.5% solid technical material caused minimal irritation. The 12.7% EC caused sensitization in the Buehler test, and the 22.9% EC caused sensitization in the local lymph node assay, with an EC3 of approximately 30%. There were no cases associated with the use of dithiopyr in the handler database. O OH OH

O O Endothall

Endothall  Endothall has complex effects on contact with plant cell membranes but also affects photosynthesis and protein synthesis (MacDonald et al., 2002). Currently, it is used principally as an aquatic, cotton, and landscape maintenance herbicide. Reported use in California for 2007 was approximately 11,000 pounds (potassium and dimethyl alkylamine salts). Physical Properties  Formula, C8H10O5; MW, 186.18; MP, 144°C; log P, 1.91; VP, 1.57E-10 mm Hg; solubility in H2O, 1.00E  05 mg/l; other solubilities (g/100 g at 20°C): acetone 7.0, benzene 0.01, dioxane 7.6, ether 0.1, isopropanol 1.7, methanol 28.0, and water 10.0 Irritation, Sensitization, and Illness Data  Three solid products containing dipotassium endothall (81.1% solid technical material, 17.9% pellets, and 63% granules) and one liquid product (28.6% AI) caused minimal irritation in the Draize assay. A 30.9% liquid product containing N, N-dimethyl alkylamine endothall salt caused corrosion. The 17.9% dipotassium endothall pellets caused sensitization in the Buehler test. There were three cases associated with endothall in the pesticide handlers database, all resulting from accidental direct contact. In one instance (case 1989-1625, Table 28.6), skin irritation was severe enough to cause a partial seconddegree burn.

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

HO

O N

N Imazethapyr

N

O

Imazethapyr  Imazethapyr is an imidazolinone and pyridine compound, inhibiting branched-chain amino acid biosynthesis. California use data for 2007 showed 55 applications, for a total of 371 pounds applied to alfalfa, beans, cotton, and sudangrass. Physical Properties  Formula C15H19N3O3; MW, 289.37; MP, 172–175°C; log P, NA; VP, NA; solubility in H2O, 1400 mg/l; other solubilities: NA Irritation, Sensitization, and Illness Data  A 97% aqueous paste (technical imazethaphyr), 70% granules, and 21.6% aqueous concentrate caused minimal irritation in the Draize test. The 22.9% formulation was a nonsensitizer in the Buehler test. There were no cases associated with imazethapyr in the handler database. OCH3

CH2CH3 C CH3

CO HN O N OCH3

CH2CH3

Isoxaben

Isoxaben  Isoxaben is a preemergent herbicide applied to the soil surface to control annual broadleaf weeds by inhibiting synthesis of cellulose necessary for plant cell walls. Reported use in California in 2007 included orchards, vineyards, greenhouse, and rights of way; there were 2721 reported applications, for a total of 22,817.3 pounds. Physical Properties  Formula, C18H24N2O4; MW, 332.44; MP, 176–179°C; log P, 3.94; VP, 4.13  10-9 mm Hg; solubility in H2O, 1.42 mg/l; other solubilities: slight solubility in organic solvents (methanol, ethyl acetate, dichloromethane, and acetonitrile) Irritation, Sensitization, and Illness Data  Technical isoxaben (approximately 93% AI) and a 75% dry flowable product caused minimal irritation in the Draize assay. The dry flowable product was also a nonsensitizer in the Buehler test. There were no cases associated with use of isoxaben in the handler database. N O

Metribuzin  Metribuzin acts by inhibition of photosynthesis (Trebst and Wietoska, 1975). California use data showed 1243 applications, for a total of 26,496.7 pounds on grains, row crops, silage, and root vegetables and for pre-plant weed control. Physical Properties  Formula C8H14N4OS; MW, 214.3; MP, 126.2°C; log P, 1.7; VP, 4.35  107 mm Hg; solubility in H2O, 1.05  103 mg/l Irritation, Sensitization, and Illness Data  Two 75% dry flowable products and a 41% flowable concentrate caused minimal irritation in the Draize assay. The 75% dry flowable and a 42.3% liquid product were negative in the Buehler test. OH Na+ O-

As

O

CH3 MSMA

Methanearsonic Acid, Monosodium Salt (MSMA) Methanearsonic acid, monosodium salt (MSMA) is an arsenical herbicide. Arsenic interacts with sulfhydryl groups on proteins and has a broad array of biological effects. However, specific interactions of MSMA with plants are not completely understood (Prukop and Savage, 1986). California use data for 2007 showed 49,878 pounds used in 428 applications on orchards, vineyards, nurseries, rights of way, and turf. Physical properties  Formula, CH4AsNaO3; MW, 161.95; MP, 130-140°C; log P, 3.10; VP, 7.8  108 mm Hg: solubility in H2O, 5.80E  05 mg/l; other solubilities: 16 g/100 ml in methanol at 25°C and 0.005 g/100 ml in hexane at 25°C Irritation and Sensitization Data  Two products containing MSMA in complex mixtures caused moderate irritation in the Draize assay. One product contained DMA salts of 3.18% 2,4-D, 0.79% dicamba, 1.6% MCPP-P, and 9.81% MSMA. The other contained DMA salts of 3.09% MCPPP, 6.21% MCPA, 1.48% dicamba, and 18.7% MSMA. The second mixture was negative in the Buehler test, as was a liquid product containing 51% MSMA. CF3 N N O

NHCH3 Cl

Norflurazon

CH3 CH3 CH3

779

N

N

NH2 Metribuzin

S CH3

Norflurazon  Norflurazon retards plant growth by inhibiting carotenoid production (Hanson and Mallory-Smith, 2000). California use data for 2007 showed 1607 applications, for a total of 77,615 pounds applied on grains, vineyards,

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orchards, row crops, outdoor plants in containers, and rights of way. Physical properties­  Formula, C12H9ClF3N3O; MW, 303.67; MP, 184°C; log P, 2.3; VP, 2.89E-08 mm Hg; solubility in H2O, 33.7 mg/l; other solubilities (g/l at 25°C): acetone 50, ethanol 142, and xylene 2.5. Irritation, Sensitization, and Illness Data  Technical solid (99.6% AI) and 5% granules caused minimal irritation in the Draize assay. The technical solid did not demonstrate sensitization in the LLNA. There was one case associated with norflurazon in the handler database – a dermatitis following contact with norflurazon on multiple occasions, despite use of protective clothing and the absence of accidental direct exposure (see Table 28.6). O O CH3 CH3

N

O

CH3

N Cl

Cl

CH3

CH3

Oxadiazon

Oxadiazon  Oxadiazon controls grasses by inhibiting photosynthesis (Routaboul et al., 2006). California 2007 use data showed 845 applications, for a total of 12,511.8564 pounds used on Bermuda grass, vineyards, nurseries, and rights of way and for landscape maintenance. Physical properties  Formula, C15H18Cl2N2O3; MW, 345.23; MP, 90°C; log P, 4.8; VP, 1.12E-07 mm Hg; solubility in H2O, 0.7 mg/l; other solubilities (g/l at 20°C): methanol and ethanol approximately 100; cyclohexane 200; acetone, isophorone, methyl ethyl ketone, and carbon tetrachloride approximately 600; and toluene, benzene, and chloroform approximately 1000. Irritation, Sensitization, and Illness Data  A product containing 50% wettable powder caused corrosion in the Draize assay. An apparently similar product containing 50% water-soluble powder caused minimal irritation. A granular product with 1.4% oxadiazon caused moderate irritation. A mixture of 1% oxadiazon and 0.5% prodiamine granules caused minimal irritation. The 50% wettable powder did not cause sensitization in the Buehler test. There were four cases associated with use of oxadiazon in the handler database, including one episode of dermatitis following direct contact (case 1988-813, Table 28.6). OCH2CH3

Cl F3C

O Oxyfluorfen

NO2

Oxyfluorfen  Oxyfluorfen is used for preemergent and postemergence control of weeds on labeled crops, inhibiting the protoporphyrinogen oxidase step in photosynthesis (Gilham and Dodge, 1987). California 2007 use data showed 45,048 applications, for a total of 644,523 12,511.8564 pounds used on orchards, grains, row crops, cotton, vineyards, nurseries, rangeland, and rights of way and for pre-plant weed control. Physical properties  Formula, C15H11ClF3NO4; MW, 361.72; MP, 84°C; BP, 358.2°C; log P, 4.73; VP, 2.48  10-7 mm Hg; solubility in H2O, 0.116 mg/l; other solubilities (g/100 g at 25°C): readily soluble in most organic solvents, such as acetone 72.5, cyclohexanone, isophorone 61.5, dimethylformamide 50, chloroform 50–55, and mesityl oxide 40–50. Irritation, sensitization, and Illness Data  The 97.1% solid technical material, two liquid products (41–42.09% AI), and three liquid mixtures (1–21% oxyfluorfen, combined with glyphosate, oxadiazon, or oryzalin) caused minimal irritation in the Draize assay. The 42.09% liquid product caused no sensitization in the GPMT. The 41% aqueous concentrate, 23% liquid, and a granular mixture of 2% oxyfluofen and 1% oxadiazon were negative in the Buehler test. There were three cases associated with oxyfluofen in the handler database, including two associated with accidental direct exposure. NHCOC2H5

Cl Cl Propanil

Propanil  Propanil is a photosynthesis inhibitor (Persch­ bachera et al., 1997) used as a rice herbicide. California use data for 2007 showed 5174 applications, for a total of 1,801,607 pounds used on approximately 378,000 acres of rice. Physical properties  Formula, C9H9Cl2NO; MW, 218.08; MP, 92°C; BP, 351°C; log P, 3.07; VP, 9.08E-06 mm Hg; solubility in H2O, 152 mg/l; other solubilities: in isopropanol and dichloromethane 200 g/l at 20°C, toluene 50– 100 g/l at 20°C, and hexane 1 g/l at 20°CC; in benzene 7  104 mg/l at 25°C, acetone 1.7  106 mg/l at 25°C, and ethanol 1.1  106 mg/l at 25°C Irritation, sensitization, and Illness Data  An 80% dry granule, a 60% dry flowable preparation, and a liquid mixture of 28.2% propanil and 0.32% bensulfuron-methyl caused minimal irritation in the Draize assay.

Chapter | 28  The Regulatory Evaluation of the Skin Effects of Pesticides

A granule containing 80% propanil and 0.62% bensulfuron and a 60% dry flowable product caused no sensitization in the Buehler test. There was one case associated with propanil in the handler database but no cases of dermatitis following accidental direct exposure. During the 1970s, cases of chloracne were associated with a manufacturing operation producing propanil (Morse et al., 1979). To this point, no chloracne has been reported with agricultural handling of propanil. O CH3 CH3CH2

S CH CH2

CH2CH2CH3 C N OCH2CH3 OH

Sethoxydim

Sethoxydim  Sethoxydim is a systemic postemergent herbicide for control of grasses. It acts by inhibiting acetyl-coenzyme A carboxylase, an enzyme important for synthesis of plastoquinones and tocopherols in plant cell membranes (Lin and Yang, 1999). California use data showed 1958 applications, for a total of 28,501 pounds on orchards, row crops, grains, uncultivated land, nurseries, and vineyards. Physical properties  Formula, C17H29NO3S; MW, 327.5; MP, 25°C; BP 90 °C at 3  10-5 mm Hg; log P, 4.38; VP, 4.55E-10 mm Hg; solubility in H2O, 25 mg/l; other solubilities: soluble in most common organic solvents including acetone, benzene, ethyl acetate, hexane, and methanol at 1 kg/kg Irritation, Sensitization, and Illness Data  The 13% liquid caused severe irritation in the Draize test but demonstrated no sensitization in the maximization test. The sethoxydim-associated case in the handler database was reported as a burn following direct accidental contact (case 1988-1253, Table 28.6).

28.3  Adjuvants Adjuvants are an important component of many applications but are the focus of less attention than the AIs. The most commonly used adjuvants are spreading and sticking agents with chemical structures similar to detergents. Some are derivatives of simple long-chain fatty acids (alkyl amino-3-aminopropane hydroxyacetate alkyl derived from coconut of fatty acids), and others are more complex synthetic molecules (e.g., alkyl aryl polyalkoxylated alcohols). Adjuvants not sold as stand-alone products are sometimes included as ingredients of formulated herbicides or insecticides (see discussion of glyphosate). Data on two adjuvants were available for this review. Neither product was

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included as an identifiable ingredient in the California illness surveillance data files. There were consequently no cases associated with this material in the handled database. Polymerized pinene is a spreading/sticking agent used for pesticide applications on turf. A Draize study conducted on this product showed mild irritation but was not carried out for a sufficient length of time to adequately characterize the category. A human repeated insult test carried out on volunteer subjects showed no evidence of sensitization. Stepan C-65 is a spray adjuvant contained in a commercial mixture with aromatic hydrocarbons (35%), phosphate ester of polyoxyalkylated fatty alcohol, and oleic acid. The mixture is labeled as a skin irritant, but no Draize study was available for review. The product was a nonsensitizer in the Buehler assay.

Conclusion The skin effects of pesticides are closely related to chemical structure and physical properties. Available tools for assessing the effects of individual compounds include predictive models based on structure–activity relationships, testing in animals, and reports regarding skin reactions in humans. The regulatory database is extensive but could be strengthened by efforts to fully explain the apparent variability in the animal test data.

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Section V

Neurotoxicology of Pesticides

(c) 2011 Elsevier Inc. All Rights Reserved.

Chapter 29

Neurotoxicology of Pesticides William Slikker, Jr. National Center for Toxicological Research, Jefferson, Arkansas

The concern over the susceptibility of the developing nervous system to pesticide exposure has continued to expand as exposure scenarios and examples of toxicity are reported. As our knowledge of the underlying biology of development and the pathways of pesticide toxicity has expanded, it has been realized that environmental factors and gene– environment interactions contribute to the expression of pesticide-induced developmental toxicity. The complexity of the developing nervous system with the differential time course for ontogeny of the various neurotransmitter receptor systems, and associated enzymes, channels, and receptors, may provide different targets for possible adverse interaction with pesticides. Although the stage of development is a major factor in the vulnerability of the nervous system, it is not currently possible to predict which stage of development, if any, will be susceptible without experimental assessment. In the postgenomic world, systems biology approaches that involve the iterative and integrative study of perturbations by chemicals of gene and protein expression that are linked firmly to toxicological outcome have been developed. Chapter 30 describes the value of systems biology to enhance the understanding of complex biological processes such as neuromodulation or neurotoxicity in the developing brain. Exposure of the developing mammal to a variety of pesticides, both alone or in combination, may perturb the endogenous neurotransmitter systems or energy metabolism/oxidative stress regulatory systems and result in enhanced neuronal cell death or dysfunction. It is proposed that continuous blockade or stimulation of various neurotransmitter receptor systems in the developing brain by pesticides that mimic or interfere with endogenous signaling pathways, receptors, or ion channels may result in long-lasting cellular dysregulation or neuronal death via apoptosis and/or necrosis. In Chapter 31, several electrophysiological targets of pesticides, including the voltage-gated sodium channel, acetylcholine receptors, and the GABAA receptors, are described. The understanding of these interactions between Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

several classes of pesticides, including the pyrethroids and organophosphate and carbamate insecticides, and these biological systems allows for the determination of the most sensitive biological component (e.g., ionic channel) but also the development of preventive strategies. Chapter 32 is devoted to understanding the differential sensitivity of the nervous system based on the developmental stage of exposure. Age-related differences in sensitivity to pesticides can be attributed to both toxicokinetic and toxicodynamic principles. Although much progress has been made, the complex nature of age-related differences in susceptibility does not allow broad-based generalities concerning the window of susceptibility even within the same class of agent. The need to define this sensitive period for each pesticide is critical for the completion of a comprehensive safety assessment. In Chapter 33, the vulnerability of the developing nervous system to organophosphate pesticides is defined in terms of cognitive functions, social and sex-related behavioral patterns, and body weight regulation. Significant behavioral alterations are reported after short-term, lowdose exposure to a variety of organophosphates during development. Because these effects were observed at doses that do not significantly inhibit acetylcholinesterase, other mechanisms in addition to inhibition of this important enzyme should to be considered. In Chapter 34, the use of the nonhuman primate model is thoroughly described for assessment of potential developmental toxicants. Because many of the current instruments that are use to assess nonhuman primates can be used in the clinical setting, cross-species extrapolation issues can be minimized with the appropriate use of this animal model. With the use of cognitive assessment tools such as the NCTR Operant Test Battery, significant adverse effects on specific brain functions have been demonstrated in the absence of effects on other, often more frequently monitored toxicological endpoints, including body weight, clinical chemistries, hematologies, and urinalyses.

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Chapter 30

A Systems Biology Approach to Assess the Impact of Pesticides on the Nervous Systema William Slikker, Jr. National Center for Toxicological Research, Jefferson, Arkansas

30.1  Introduction Systems biology/toxicology involves the iterative and integrative study of perturbations by chemicals and other stressors of gene and protein expression that are linked firmly to toxicological outcome. In this chapter, the value of systems biology to enhance the understanding of complex biological processes such as neuromodulation or neurotoxicity in the developing brain is explored. Exposure of the developing mammal to a variety of pesticides, both alone or in combination, may perturb the endogenous neurotransmitter systems or energy metabolism/oxidative stress regulatory systems and result in enhanced neuronal cell death or dysfunction. It is proposed that continuous blockade or stimulation of various neurotransmitter receptor systems in the developing brain by pesticides that mimic or interfere with endogenous signaling pathways, receptors or ion channels may result in long-lasting cellular dysregulation or neuronal death via apoptosis and/or necrosis. Systems biology has been defined as the iterative and integrative study of biological systems as they respond to perturbations (Auffray et al., 2003; Hood and Galas, 2003; Ideker et al., 2001). In this chapter, systems biology is explored as an approach to enhance the understanding of complex biological processes such as pesticide-induced neuronal modulation or neurodegeneration in the developing nervous system. High throughput or high density data molecular biology approaches including genomics, proteomics, and

a Disclaimer: The views presented in this overview do not necessarily reflect those of the U.S. FDA.

Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

metabolomics provide the fundamental data necessary for the building blocks of systems biology. As these databases grow and become linked together as integrated modules, they will provide the intermediate components necessary for the systems biology approach. It is the appropriate placement of these biological modules or pathways into a proposed mechanistic flow scheme, thus allowing for the development of integrated computational models, that is the overall goal. However, the development of these mathematical models often lags behind the initial definition of the system and often remains to be accomplished. For toxicology, it is essential that quantitative correlations of exposure and response (i.e., dose, time intervals, and outcome) be integrated into the computational model (Henry, 2003). In addition to knowledge about the proximate toxicant and its mechanism of action, the primary toxi­ cological effect or phenotypic anchor must also be utilized (Waters et al., 2003b). At the systems biology level, quantitative simulations can be conducted and predictions of the model can be tested. The outcome of these iterations is systematically incorporated back into the model to improve its design and refine its predictive capabilities. The interconnectivity of a system at this level determines its state and extends its predictive power (Jazwinski, 2002). The goal of systems biology is to predict the functional outcomes of component-to-component relationships using computational models that allow for the directional and quantitative description of the complete organism in response to environmental perturbations (Waters et al., 2003a). Systems biology approaches can also be used as effective tools for dissecting the mechanisms underlying toxicological phenomena associated with exposure to toxicants. It is the development of predictive models that 793

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integrate responses across different organizational levels that is the focus of this chapter.

30.2  Pesticides and developmental toxicity Recently described studies of pesticides will be used to exemplify the advantages of the application of the systems biology approach. The developing nervous system may be more or less susceptible to neurotoxic insult depending on the stage of development. Because of the complexity and temporal features of the manifestations of developmental neurotoxicity, this area of toxicology can benefit from a systems biology approach. The main purpose of this chapter is to outline the application of the systems bio­ logy approach to the issue of toxicity produced by pesticides. It is proposed that the administration of pesticides during critical developmental periods will result in a doserelated increase in toxicity including neurotoxicity (loss of neurons or neuronal dysfunction) by a mechanism that involves altered gene expression, protein elaboration, or endogenous metabolite modification. In order to predict if pesticide-induced toxicity in the developing mammal has clinical relevance, agents can be examined in a variety of in vivo and in vitro biological models that closely mimic the pediatric population. The four steps of a systems biology approach reported by Leroy Hood’s group (Auffray et al., 2003) will be discussed in this context. First, available information on the biological system of interest should be described and a preliminary model of how the system functions should be formulated. Second, where possible, the genes and proteins expressed in the described pathways should be defined. However, information about genetic perturbations of the system is generally not available. Third, kinetic experiments providing information across important stages of development should be considered. Fourth, various global datasets should be integrated to determine if they support the model. Discrepancies should be identified and ­hypotheses-driven studies should be conducted in order to address them. Thus, data generated via iteration of the third and fourth steps should be used to reformulate the model in light of new data. Although mathematical modeling is an ultimate goal of the systems biology approach, it is often, at this time, not achievable.

30.3  Developmental neurotoxicity The concern over the susceptibility of the developing nervous system to toxic insult has continued to expand as exposure scenarios and examples of toxicity are reported (Bellinger, 2007; Daston et al., 2004; Ginsberg �� et al., 2004��� ). Although the range of central nervous system anomalies,

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including learning disabilities through mental retardation, has been reported in the range from 10% to 20%, it is known that not all nervous system dysfunction is caused by toxic exposure (Lipkin, 1991). Genetic syndromes and unknown causes are a component of this overall rate of developmental toxi­city. While it is likely that some of the risk factors for these and other behavioral/functional disorders are genetic, it is almost certain that environmental factors and gene– environment interactions contribute to the expression of these clinical entities (phenotypes) (Rodier et al., 1994). One major change in the study of developmental neurotoxicity is the scientific approach used to address the issue. In the post-genomic world, the application of a systems biology approach to understand developmental neurotoxicity is possible. Because of the complexity and temporal features of the manifestations of developmental neurotoxicity, no area of toxicology can benefit more from the systematic application of the systems biology/systems toxicology approach. Neurotoxicity may be defined as any adverse effect on the structure or function of the central and/or peripheral nervous system by a biological, chemical, or physical agent that diminishes the ability of an organism to survive, reproduce, or adapt to its environment. Neurotoxic effects may be permanent or reversible, produced by neuropharmacological or neurodegenerative properties of a toxicant, or the result of direct or indirect actions on the nervous system (Slikker, 1991). These effects can often be measured by neurobiological, neurophysiological, neuropathological or behavioral techniques. Extrapolation across species is feasible but must take into account the relative ontogeny of the nervous system among species. Insults to the nervous system may take various forms and may be quite subtle (Anger, 1986). Although its manifestations may change with age, neurotoxicity may occur at any time in the lifecycle from gestation through senescence. The developing nervous system may be more or less susceptible to neurotoxic insult depending on the stage of maturity.

30.4  Examples from the current literature One general example is provided by Gohlke et al. (2009) and focuses on a broad range of toxicants. According to systems theory, although individual genes or environmental factors may be a critical component in the pathogenesis of a particular complex disease, the adverse effect on phenotype is often the result of modulation of underlying pathways of which that particular gene/environmental factor is a part. Gohlke and coworkers have integrated genecentered knowledge from epidemiological and mechanistic environmental research in an attempt to discover the interplay between genetic and environmental mediators of phenotype at the pathway level. They provided a higher order

Chapter | 30  A Systems Biology Approach to Assess the Impact of Pesticides on the Nervous System

structure of pathway interconnectivity to build hypotheses of disease progression based on clusters of pathways defining phenotypes. The authors suggest that the methods and findings allow a number of new hypotheses that can be explored regarding the genetic and environmental factors governing human disease. The results suggest retinol metabolism, Jak-STAT signaling, Toll-like receptor signaling, and adipocytokine signaling are key pathways that should be prioritized targets for high-throughput screening currently being implemented to improve toxicity testing (NTP, 2004; NRC, 2007). For example, analysis of the metabolic syndrome sub-network highlights the need for further epidemiological and mechanistic analyses of several compounds for their potential modulation of metabolic syndrome phenotypes, including plastic derivatives, synthetic and natural retinoids, antipsychotic medications, and pyrethrin pesticides. The search for the critical biological assays to identify appropriate cellular toxicity pathways for interrogation using biochemical- and cell-based high-throughput screens is underway (Martin et al., 2009). High-throughput assays already performed at the National Center for Genomic Research (NCGC) include those to assess: (1) cytotoxi­ city and activation of caspases in a number of human and rodent cell types; (2) upregulation of p53; (3) agonist/ antagonist activity for a number of nuclear receptors; and (4) differential cytotoxicity in several cell lines associated with an inability to repair various classes of DNA damage. Other assays under consideration include those for a variety of physiologically important molecular pathways (e.g., cellular stress responses) as well as methods for integrating human and rodent hepatic metabolic activation into reporter gene assays. Based on the results obtained for hundreds of chemicals including pesticides, these researchers plan to construct test batteries useful for identifying hazard for humans and for prioritizing chemicals including pesticides for further, more in-depth evaluation (Kavlock et al., 2009). Similar pathway-based assessment approaches are being reported for developmental toxicity and include accounting for manifestations of direct (mechanism-based) developmental toxicity with or without indirect (maternalmediated) effects (Knudsen et al., 2009). It is envisioned that data from alternative methods and high-throughput in vitro assays that enable pathway-based risk assessment may increase confidence in testing strategies while limiting required animal testing (Bremer et al., 2007; NRC, 2007). The authors suggest that toxicity reference databases provide a novel data model for relational assessment of source data from guideline (in vivo) prenatal developmental toxi­ city studies to anchor cross-scale modeling and predictive understanding of developmental processes and toxicities (Knudsen and Kavlock, 2008). Although the application of systems biology approaches has been applied to the study of developmental neurotoxicology (Slikker et al., 2005, 2006), the

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systematic application of a systems approach to study pesticide-induced developmental neurotoxicity has not been routinely reported. The encompassing nature of the systems biology approach and its utility in uncovering pathways and networks leading to toxicity can be applied to the study of pesticides during development. One example is the organophosphates, including chlorpyrifos, diazinon, and parathion, that account for up to 50% of all insecticide application worldwide (Casida and Quistad, 2004). Organophosphates are well known for their systemic toxi­city based on their ability to inhibit cholinesterase, but it has been reported that their developmental toxicity may be attributed to other mechanisms including cell signaling cascades governing homeostatic regulation and cellular differentiation (Barone et al., 2000; Slotkin, 2004). The effects on the developing nervous system are extensive and include alterations of cell replication and differentiation, interference with axonogenesis and synaptogenesis, and impairment of several neurotransmitter systems beyond the acetylcholine system including the serotonin (5HT) system (Slotkin et al., 2008). In addition to the broad range of effects on the developing nervous system, organophosphate pesticides have been reported to have other lasting effects culminating in a metabolic pattern characteristic of dyslipidemia and prediabetes (Lassiter et al., 2008; Slotkin et al., 2005). When developing rats were exposed to chlorpyrifos (1 mg/kg, postnatal days 1–4, a dosing regimen below the threshold for systemic toxicity), and tested in adulthood, the researchers observed gender-selective elevations in plasma cholesterol and triglycerides in male offspring. The authors conclude that low-level organophosphate exposure results in a metabolic pattern for plasma lipids and insulin that resembles the major adult risk factors for atherosclerosis and type 2 diabetes mellitus (Lassiter et al., 2008; Slotkin et al., 2005). The interpretation of these developmental exposure results following exposure to organophosphates, indicating the activation of numerous and associated pathways of toxicity, calls for the use of a systems biology approach. The complexity and interrelatedness of the many altered pathways requires the development of a model reflecting the selective but numerous gene expression, protein production, and metabolic modulations. Armed with this overview and scope of the biological perturbations, the researchers can begin to define the leading hypotheses worth further study and define the primary targets, refine critical biomarkers, and develop prevention strategies. Another example of the complex biological effects of pesticides that could be clarified with a systems biology approach is the role of paraoxonase 1 (PON1) polymorphisms in organophosphate neurotoxicity (Costa et al., 2003). This enzyme was initially characterized as an organophosphate hydrolase and its name is derived from one of the most commonly used pesticides, paraoxon.

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Because only organophosphates with a PO moiety interact with acetylcholinesterase to enhance acetylcholine levels that produce the characteristic overstimulation of nicotinic and muscarinic receptors in the central and peripheral nervous system, cholinergic, A-esterases, including PON1, can hydrolyze and thus detoxify a number of organophosphates (Costa and and Furlong, 2002; Davies et al., 1996). PON1 is also a high-density lipoprotein (HDL)-associated serum enzyme and has the primary role to protect lowdensity lipoproteins (LDL) from oxidative modifications (Mackness et al., 1991). More recent studies have implicated PON1 in the metabolism of pharmaceutical drugs as well as its prominent role in lipid metabolism and impact in cardiovascular disease and atherosclerosis (Costa et al., 2003). Polymorphisms in both the coding and regulatory regions of the PON1 gene help determine an individual’s PON1 plasma activity and thus have been postulated as predictors of individual sensitivity to organophosphate pesticides, certain pharmaceuticals, and cardiovascular risk (Costa et al., 2003). The capacity of these A-esterases to detoxify organophosphates is also reported to be dependent on the maturational stage of the organism highlighting the importance of enzyme ontogeny in pesticide safety assessment (Costa et al., 1990; Karanth and Pope, 2000). Here too, the application of the systems biology approach, describing the full extent of PON1 gene expression, protein production and metabolites and its polymorphic variation, would help clarify the fuller role of this complex and multi-impact serum enzyme that affects both the nervous and cardiovascular systems.

Conclusions The success of the systems biology approach to solve toxicological problems lies in the establishment of crossdisciplinary teams of scientists including toxicologists, pathologists, molecular biologists, mathematicians, statisticians, computational modelers, and risk assessors. The integration of rapidly growing biological databases, including models of cells, tissues and organs, with the use of powerful computing systems and algorithms is necessary (Noble, 2003). These interdisciplinary scientists are conducting systematic experiments that account for small variations in a large number of model components in order to determine the overall functioning of the biological system (Auffray et al., 2003). High-density data or high-throughput molecular bio­ logy approaches including genomics, proteomics and metabonomics are providing the fundamental data necessary for the building blocks of a systems biology approach to predict developmental neurotoxicity. As these databases grow and become linked together as integrated modules or toxicity pathways, they will provide the intermediate components necessary for the systems biology approach.

It is the appropriate placement of these biological modules or toxi­city pathways into a wiring diagram, allowing the development of an integrated computational model, which remains to be accomplished (Auffray et al., 2003). It is at this level that connectivity of the system determines its state and the whole becomes greater than the sum of its parts (Jazwinski, 2002). For toxicology, it is essential that quantitative correlations of exposure, dose and outcome be integrated into the computational model (Henry, 2003). In addition to knowledge of the proximate toxicant and its mechanism of action, the primary toxicological effect or phenotypic anchor must be incorporated into the model (Waters et al., 2003b). At this systems biology level, quantitative simulations can be conducted and predictions of the model output can be tested. The quantitative outcome of these iterative experiments is systematically incorporated back into the model to improve its design and refine its predictive capabilities. Examples of emerging systems biology applications that have been reported include the integration of genecentered knowledge from epidemiological and mechanistic research in an attempt to discover the interplay between genetic and environmental mediators of phenotype at the pathway level. Although the global integrative processes are generally depicted in graphic form, mathematically based models are essential for the full potential of systems biology to be achieved. Powerful mathematical approaches have been used to describe quantitatively physiologically based pharmacokinetic models with pharmacodynamic components (Doerge et al., 2008; Timchalk et al., 2002), but these existing frameworks based on simultaneously solved differential equations have yet to be applied routinely to systems biology assessment approaches for pesticides.

References Anger, W. K. (1986). Worker exposures. In “Neurobehavioral Toxicology,” (Z. Annau, ed.), pp. 331–347. Johns Hopkins Press, Baltimore, MD. Auffray, C., Imbeaud, S., Roux-Rouquie, M., and Hood, L. (2003). From functional genomics to systems biology: concepts and practices. C.R. Biol. 326, 879–892. Barone, S. Jr., Das, K. P., Lassiter, T. L., and White, L. D. (2000). Vulnerable processes of nervous system development: a review of markers and methods. Neurotoxicology 21(1–2), 15–36. Bellinger, D. C. (2007). Children’s cognitive health: the influence of environmental chemical exposures. Altern. Ther. Health Med. 13(2)S, 140–144. Bremer, S., Pellizze, R. C., Hoffmann, S., Seidle, T., and Hartung, T. (2007). The development of new concepts for assessing reproductive toxicity applicable to large scale toxicological programmes. Curr. Pharm. Des. 13, 3047–3058. Casida, J. E., and Quistad, G. B. (2004). Organophosphate toxicology: safety aspects of nonacetylcholinesterase secondary targets. Chem. Res. Toxicol. 17, 983–998.

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A Systems Biology Approach to Assess the Impact of Pesticides on the Nervous System

Costa, L. G., and Furlong, C. E. (2002). Perspectives in PON research in paraoxonase (PON1). In “Health and Disease: Basic and Clinical Aspects” (L. G. Costa and C. E. Furlong, eds.), pp. 165–183. Kluwer Academic, Norwell, MA. Costa, L. G., McDonald, B. E., Murphy, D. S., Omenn, G. S., Richter, R. J., Motulsky, A. G., and Furlong, C. E. (1990). Serum paraoxonase and its influence on paraoxon and chlorpyrifos-oxon toxicity in rats. Toxicol. Appl. Pharmacol. 103, 66–76. Costa, L. G., Cole, T. B., Jarvik, G. P., and Furlong, C. E. (2003). Functional genomic of the paraoxonase (PON1) polymorphisms: effects on pesticide sensitivity, cardiovascular disease, and drug metabolism. Annu. Rev. Med. 54, 371–392. Daston, G., Faustman, E., Ginsberg, G., Fenner-Crisp, P., Olin, S., Sonawane, B., Bruckner, J., Breslin, W., and McLaughlin, T. J. (2004). A framework for assessing risks to children from exposure to environmental agents. Environ. Health Perspect. 112, 238–256. Davies, H. G., Richter, R. J., Keifer, M., Broomfield, C. A., Sowalla, J., and Furlong, C. E. (1996). The effect of the human serum paraoxonase polymorphism is reversed with diazoxon, soman and sarin. Nat. Genet. 14, 334–336. Doerge, D. R., Young, J. F., Chen, J. J., Dinovi, M. J., and Henry, S. H. (2008). Using dietary exposure and physiologically based pharmacokinetic/pharmacodynamic modeling in human risk extrapolations for acrylamide toxicity. J. Agric. Food Chem. 56, 6031–6038. Ginsberg, G., Slikker, W. Jr., Bruckner, J., and Sonawane, B. (2004). Incorporating children’s toxicokinetics into a risk framework. Environ. Health Perspect. 112, 272–283. Gohlke, J. M., Thomas, R., Zhang, Y., Rosenstein, M. C., Davis, A. P., Murphy, C., Becker, K. G., Mattingly, C. J., and Portier, C. J. (2009). Genetic and environmental pathways to complex diseases. BMC Syst. Biol. 3, 46. Henry, C. J. (2003). Evolution of toxicology for risk assessment. Int. J. Toxicol. 22, 3–7. Hood, L., and Galas, D. J. (2003). The digital code of DNA. Nature 421, 444–448. Ideker, T., Galitski, T., and Hood, L. (2001). A new approach to decoding life: systems biology. Annu. Rev. Genom. Hum. Genet. 2, 343–372. Jazwinski, S. M. (2002). Biological aging research today: potential, peeves, and problems. Exp. Gerontol. 37, 1141–1146. Karanth, S., and Pope, C. (2000). Carboxylesterase and A-esterase activities during maturation and aging: relationship to the toxicity of chlorpyrifos and parathion in rats. Toxicol. Sci. 58, 282–289. Kavlock, R. J., Austin, C. P., and Tice, R. R. (2009). Toxicity testing in the 21st century: implications for human health risk assessment commentary. Risk Analysis 29, 485–487. Knudsen, T. B., and Kavlock, R. J. (2008). Comparative bioinformatics and computational toxicology. In “Developmental Toxicology.” Vol. 3, Target Organ Toxicology Series (B. Abbott and D. Hansen, eds.), pp. 311–360. Taylor and Francis, New York. Knudsen, T. B., Martin, M. T., Kavlock, R. J., Judson, R. S., Dix, D. J., and Singh, A. V. (2009). Profiling the activity of environmental chemicals in prenatal developmental toxicity studies using the U.S. EPA’s ToxRefDB. Reprod. Toxicol. 28, 209–219. Lassiter, T. L., Ryde, I. T., Mackillop, E. A., Brown, K. K., Levin, E. D., Seidler, F. J., and Slotkin, T. A. (2008). Exposure of neonatal rats to parathion elicits sex-selective reprogramming of metabolism and

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alters the response to a high-fat diet in adulthood. Environ. Health Perspect. 116, 1456–1462. Lipkin, P. H. (1991). Epidemiology of developmental disabilities. In “Developmental Disabilities in Infancy and Childhood” (A. J. Capute and P. J. Accardo, eds.), pp. 43–67. Brookes, Baltimore, MD. Mackness, M. I., Arrol, S., and Durrington, P. N. (1991). Paraoxonase prevents accumulation of lipoperoxides in low-density lipoprotein. FEBS Lett. 286, 152–154. Martin, M. T., Mendez, E., Corum, D. G., Judson, R. S., Kavlock, R. J., Daniel, M., Rotroff, D. M., and Dix, D. J. (2009). Profiling the reproductive toxicity of chemicals from multigeneration studies in the toxicity reference database (ToxRefDB). Toxicol. Sci. 110, 181–190. Noble, D. (2003). The future: putting humpty-dumpty together again. Biochem. Soc. Trans. 31, 156–158. NRC (2007). Toxicity testing in the 21st century, a vision and a strategy. In “Council NRC” p. 196. National Academies Press, Washington, DC. NTP (2004) “A National Toxicology Program for the 21st Century: A Roadmap to Achieve the NTP Vision.” (National Toxicology Program/National Institute of Environmental Health Sciences, eds.), Research Triangle Park, NC. Rodier, P. M., Cohen, I. R., and Buelke-Sam, J. (1994). Neuroendocrine manifestations of CNS insult. In “Developmental Toxicology” (C. A. Kimmel and J. Buelke-Sam, eds.), 2nd ed., pp. 65–92. Raven Press, New York. Slikker, W. Jr. (1991). Biomarkers of neurotoxicity: an overview. Biomed. Environ. Sci. 4, 192–196. Slikker, W. Jr., Xu, Z., and Wang, C. (2005). Application of a systems biology approach to developmental neurotoxicology reproductive toxicology. J. Reprod. Toxicol. 19, 305–319. Slikker, W. Jr., Paule, M. G., Wright, L. K. M., Patterson, T. A., and Wang, C. (2006). Systems biology approaches for toxicology. J. Appl. Toxicol. 27, 201–217. Slotkin, T. A. (2004). Cholinergic systems in brain development and disruption by neurotoxicants: nicotine, environmental tobacco smoke, organophosphates. Toxicol. Appl. Pharmacol. 198, 132–151. Slotkin, T. A., Brown, K. K., and Seidler, F. J. (2005). Developmental exposure of rats to chlorpyrifos elicits sex-selective hyperlipidemia and hyperinsulinemia in adulthood. Environ. Health Perspect. 113, 1291–1294. Slotkin, T. A., Ryde, I. T., Levin, E. D., and Seidler, F. J. (2008). Developmental neurotoxicity of low dose diazinon exposure of neonatal rats: effects on serotonin systems in adolescence and adulthood. Brain Res. Bull. 75, 640–647. Timchalk, C., Nolan, R. J., Mendrala, A. L., Dittenber, D. A., Brzak, K. A., and Mattsson, J. L. (2002). A physiologically based pharmacokinetic and pharmacodynamic (PBPK/PD) model for the organophosphate insecticide chlorpyrifos in rats and humans. Toxicol. Sci. 66, 34–53. Waters, M., Boorman, G., Bushel, P., Cunningham, M., Irwin, R., Merrick, A., Olden, K., Paules, R., Selkirk, J., Stasiewicz, S., Weis, B., Van Houten, B., Walker, N., and Tennant, R. (2003b). Systems toxicology and the chemical effects in biological systems (CEBS) knowledge base. EHP Toxicogenomics 111, 15–28. Waters, M. D., Olden, K., and Tennant, R. W. (2003a). Toxicogenomic approach for assessing toxicant-related disease. Mutat. Res. 544, 415–424.

Chapter 31

Neurophysiological Effects of Insecticides Toshio Narahashi Northwestern University Medical School

The latter half of the 20th century has witnessed a considerable advance in our knowledge concerning the mechanisms of action of insecticides. This was due mostly to impressive developments of newer, mostly synthetic, insecticides, and rapid progress in various technologies in the field of biomedical sciences. Among various areas of the insecticide mechanism of action, studies of their metabolism were among the earliest developments, starting in the 1950s. However, it was not until the 1960s that studies of the cellular or physiological mechanism of action of insecticides became widespread. More recently, applications of molecular biology and genetics techniques have made it possible to identify the molecular species that are responsible for the toxic action of insecticides, particularly those related to the target resistance of insects to insecticides. Most insecticides are neuropoisons, but their target sites are rather limited. For example, voltage-gated sodium channels are the major target of pyrethroids and DDT; GABAA receptors are attacked by cyclodienes, hexachlorocyclohexane (HCH), and fipronil; neuronal nicotinic acetylcholine (nnACh) receptors are the target of nicotine, and nitromethylene and nitroimine hete-rocycles (e.g., imidacloprid). Organophosphate and carbamate insecticides inhibit acetylcholinesterase. This chapter covers the neurophysiological mechanisms of action of various insecticides. However, since a large number of review articles have already been published, emphasis will be placed on recent developments in the field. Readers are encouraged to refer to these review articles, each of which discusses similar issues from somewhat different points of view. These articles, though not limited to, are as follows: Narahashi (1971, 1976, 1985, 1988, 1989, 1992, 1996), Narahashi et al. (1995, 1998), Ruigt (1984), Soderlund and Bloomquist (1989), Vijverberg and van den Bercken (1990), Salgado (1999), Clark (1997), Bloomquist (1996), and Casida and Quistad (1998).

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31.1  Pyrethroids and DDT 31.1.1  Sodium Channel Modulation Despite apparent differences in chemical structure, pyrethroids and DDT exert similar actions on the nervous system through modulation of the function of voltage-gated sodium channels. Pyrethroids may be divided into two groups: type I pyrethroids lack a cyano group in the  position, and their symptoms of poisoning are characterized by hyperexcitation, ataxia, convulsions, and paralysis; type II pyrethroids have an  cyano group, and cause hypersensitivity, choreoathetosis, tremors, and paralysis. At the level of nerve function, type I pyrethroids tend to produce repetitive action potentials as a result of the increase in depolarizing after-potential, whereas type II pyrethroids tend to cause membrane depolarization leading to discharges from sensory neurons. These apparent differences in nerve function alteration between the two types of pyrethroids can be ascribed to differences in modification of sodium channel kinetics. DDT has many features in common with type I pyrethroids with respect to the mechanism of action on the sodium channel. Changes in Sodium Channel Gating Kinetics Depolarizing after-potential is gradually increased after application of type I pyrethroids such as tetramethrin and allethrin, and reaches the threshold membrane potential for generation of action potentials (Lund and Narahashi, 1981a, b; Narahashi, 1962; Vijverberg et al., 1982). The mechanism by which the depolarizing after-potential is increased can best be studied by the voltage clamp technique (Fig. 31.1). The tail current upon termination of a depolarizing pulse was greatly increased and prolonged in the presence of pyrethroid. Type II pyrethroids such as deltamethrin and fenvalerate caused much greater prolongation of sodium currents during and upon termination of a depolarizing pulse

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(a)

(b)

0.6 mA/cm2 2 ms Figure 31.1  Effects of 1 M ()-trans allethrin on the sodium current of a squid giant axon. The membrane was step depolarized to 20 mV from a holding potential of 100 mV in K-free external and internal prefusates. In the control (a), the peak transient sodium current is followed by a small slow current during a depolarizing step, and the tail sodium current upon step repolarization decays quickly. After application of allethrin (b), the peak transient sodium current remains unchanged, but the slow current and tail current are increased in amplitude and the latter decays very slowly. From Narahashi (1984).

than type I pyrethroids (Brown and Narahashi, 1987, 1992; Ogata et al., 1988; Salgado et al., 1989; Song et al., 1996; Tabarean and Narahashi, 1998). Cockroach neurons cultured from the brain of 21-day-old embryos did not express sodium channel activity, yet deltamethrin unveiled “silent” sodium channels which were partly blocked by tetrodotoxin (TTX) (Amar and Pichon, 1992). Pyrefhroid modulation of individual sodium channels was studied by single-channel patch clamp techniques using neuroblastoma cells. While normal sodium channels opened for a few milliseconds at the beginning of a depolarizing pulse, channels exposed to pyrethroid opened for a very long period of time often extending a few seconds and with a long delay from the beginning of the depolarizing pulse (Fig. 31.2) (Chinn and Narahashi, 1986; Holloway et al., 1989; Yamamoto et al., 1983). In the presence of pyrethroid, sodium channels often remained open after termination of the depolarizing pulse reflecting the whole-cell tail current. These observations have led to the conclusion that the kinetics of both activation and inactivation gates are slowed and the gates tend to be stuck at the open or closed position (Chinn and Narahashi, 1986; Vijverberg et al., 1982). As expected from these results, the gating currents associated with both opening and closing of the sodium channel were inhibited by pyrethroid (Salgado and Narahashi, 1993). Extremely prolonged sodium channel openings (up to several seconds) were also observed in cockroach neurons in culture in the presence of deltamethrin (Amar and Pichon, 1992). State Dependency of Sodium Channel Modification A drug could bind to a channel at its closed state, its open state, or both states. This is an important aspect of drug– channel interaction. Extensive studies along this line have led to the conclusion that pyrethroids modify the sodium channel function at its closed state but the open channel has a higher affinity for pyrethroids. Thus, pyrethroids act

–30 mV

–100 mV

1 pA 10 ms (a) 1 2 1 pA 3 4 5 6 7 8 –30 mV –100 mV

140 ms

3s

(b) Figure 31.2  Deltamethrin prolongation of single sodium channel currents recorded from a neuroblastoma cell (N1E-115). (a) Currents from a cell before drug treatment in response to 140-msec depolarizing steps from a holding potential of 100 mV to 30 mV with a 3-sec interpulse interval. Records were taken at a rate of 100 sec per point. (b) Currents after exposure to 10 M deltamethrin. The membrane patch was depolarized for 3140 msec from a holding potential of 100 mV to 30 mV. The interpulse interval was 3 sec. The time scale changed during the voltage step as indicated in the figure. During the first 140 msec, records were taken at a rate of 100 sec per point, and after the vertical line, records were taken at a rate of 10 msec per point. From Chinn and Narahashi (1986).

Chapter | 31  Neurophysiological Effects of Insecticides

on the closed sodium channel, and opening further recruits modified channels (Brown and Narahashi, 1992; de Weille et al., 1988; Ginsburg and Narahashi, 1999; Holloway et al., 1989). It should be noted that the ratio of open sodium channel modification to closed sodium channel modification varies considerably in different preparations. For example, a large fraction of pyrethroid modification occurred in the closed channel state in squid giant axons (de Weille et al., 1988), in rat dorsal root ganglion neurons (Ginsburg and Narahashi, 1999), and in mouse neuroblastoma cells (Holloway et al., 1989), whereas pyrethroid modification occurred largely in the open channel state in frog muscle fibers (Leibowitz et al., 1986). Open Channel Properties While passing through an open sodium channel, the permeating cation must cross barriers, and temporarily binds to sites inside the channel. Thus, open sodium channels are not only permeable to but also blocked by various monovalent and divalent cations to a varying extent. The permeability ratios in squid axons for Na:Li:ammonium: guanidine:formamidine were 1:1.19:0.21:0.28:0.20 for the normal sodium channel, and 1:1.18:0.29:0.29:0.25 for the channel modified by tetramethrin (Yamamoto et al., 1986). It is concluded that pyrethroid does not alter the permeability properties of open sodium channel. Site of Action of Pyrethroids in the Sodium Channel A variety of experimental approaches have been taken to determine the site of action of pyrethroids in the sodium channel. Pyrethroids have been shown to bind to a site different from any other known sites for various toxins and chemicals. n-Octyl-guanidine blocked the sodium channel by entering from inside the membrane when the gates are open (Kirsch et al., 1980) in a manner similar to that of local anesthetics (Courtney, 1975; Hille, 1977; Strichartz, 1973; Yeh, 1978, 1980), pancuronium (Yeh and Narahashi, 1977), 9-aminoacridine (Yeh, 1979), and strychnine (Shapiro, 1977). The octylguanidine binding site was not the site for pyrethroids as they did not interact with each other (de Weille et al., 1988). Batrachotoxin (BTX) and grayanotoxin (GTX) slow the kinetics of the activation and inactivation gates of the sodium channel and shift their voltage dependence in the hyperpolarizing direction, resulting in slow and prolonged sodium current and membrane depolarization (Albuquerque et al., 1971; Khodorov, 1985; Khodorov et al., 1976; Narahashi et al., 1971; Narahashi and Seyama, 1974; Seyama and Narahashi, 1973, 1981; Tanguy and Yeh, 1991). BTX and GTX are also known to bind to site 2 of the sodium channel (Catterall, 1992). Tetramethrin action was not modified by either BTX (Tanguy and Narahashi, unpublished) or GTX (Takeda and Narahashi, 1988). Therefore, pyrethroids bind to a site other than site 2. TTX selectively blocks the sodium channel (Narahashi

801

et al., 1964) through binding to site 1 (Catterall, 1992). TTX blocked the tetramethrin-modified sodium channel in a noncompetitive manner, indicating that tetramethrin did not bind to the TTX site (site 1) (Lund and Narahashi, 1982). A recent study has shown that pyrethroids modify the  subunit of the sodium channel expressed in Chinese hamster ovary cells through binding to a site other than any other known binding sites for various toxins and chemicals (Trainer et al., 1997). Binding of [3H]batrachotoxinin A-20--­benzoate 3 ([ H]BTX-B) to mouse brain sodium channels was modified by pyrethroids and DDT (Rubin et al., 1993). Although deltamethrin and the 2S stereoisomers of fenvalerate enhanced [3H]BTX-B binding, nontoxic isomers inhibited the binding or caused no effect. DDT and its analogs and metabolites enhanced the binding. However, toxic type I pyrethroids enhanced, inhibited, or had no effect on the binding, and the effects were not correlated with toxicity. These data illustrate a limitation in the use of this assay as a screen for neurotoxicity (Rubin et al., 1993). The role of  and 1 subunits of rat brain IIa sodium channel in pyrethroid action was studied using Xenopus oocyte expression and voltage clamp techniques (Smith and Soderlund, 1998). In both the  and the  plus 1 subunits expressed in oocytes, cypermethrin caused prolonged tail sodium currents. However, the cypermethrin affinity was 20 times higher in the  plus 1 combination than in the  subunit alone. Differential Pyrethroid Sensitivity to TTX-Sensitive and TTX-Resistant Sodium Channels Most sodium channels in the nervous system are highly sensitive to TTX block, with an IC50 in the range of nanomolar concentrations. By contrast, cardiac sodium channels are less sensitive to TTX, with an IC50 on the order of micromolar concentrations. During the past several years, TTX-­resistant (TTX-R) sodium channels in the nerve have received much attention, partly because some of these channels in mammalian dorsal root ganglia (DRG) are related to pain sensation, opening the door for the possible development of drugs that selectively block TTX-R sodium channels as useful analgesics. Although the initial discovery of TTX-R sodium channels in DRG was made almost 20 years ago by Kostyuk et al. (1981), it was not until the early 1990s that their significance received much attention after being revisited by Roy and Narahashi (1992). The IC50 for TTX-R sodium channels was about 100 M, a value 100,000 times higher than that for TTX-sensitive (TTX-S) sodium channels. Analyses of TTX-R as well as TTX-S sodium channels have been performed extensively not only for their physiology and biophysics (Elliott and Elliott, 1993; Ogata and Tatebayashi, 1993), but also for their molecular structures (Akopian et al., 1996; Sangameswaran et al., 1997). Significance of TTX-R sodium channels in insecticide toxicology has been demonstrated for pyrethroids. TTX-R

Hayes’ Handbook of Pesticide Toxicology

802

sodium channels of rat DRG neurons were more sensitive to pyrethroid modulation than TTX-S sodium channels of DRG neurons (Ginsburg and Narahashi, 1993; Tatebayashi and Narahashi, 1994). An example of such a patch clamp experiment is shown in Fig. 31.3. Although TTX-S sodium channel current during a depolarizing pulse was only slightly affected by 1 M tetramethrin (Fig. 31.3a), TTX-R sodium channel current underwent drastic changes including the appearance of a large tail current upon termination of the depolarizing step (Fig. 31.3b). Similar differential sensitivity to pyrethroids was also found between insect and mammalian sodium channels. Currents were recorded from Xenopus oocytes expressing para sodium channel  subunit from Drosophila and rat brain type IIA sodium channels (Warmke et al., 1997). Permefhrin was over 100 times more potent in modulating sodium currents of para sodium channels than those of brain IIA sodium channels. The differential sodium channel sensitivity is one of the crucial factors that account for the selective toxicity of pyrethroids, as will be discussed later. Amplification of Pyrethroid Toxicity from Sodium Channels to Animals An early study by Lund and Narahashi (1982) using squid giant axons suggested that only a very small fraction of the sodium channel

0 –110

Control

Tetramethrin (1 µM)

4 nA 10 msec

(a) 0 –90

Control Tetramethrin (1 µM)

4 nA

10 msec (b) Figure 31.3  Effects of tetramethrin on tetrodotoxin (TTX)-sensitive sodium current (a) and TTX-resistant sodium current (b) in rat dorsal root ganglion neurons. A step depolarization to 0 mV was applied from a holding potential of 110 mV (a) or 90 mV (b) in control and in the presence of 1 M tetramethrin. From Tatebayashi and Narahashi (1994).

population needed to be modified by pyrethroids to cause repetitive discharges. This was based on the calculation of the percentage of sodium channels needed to increase the depolarizing after-potential to the level of threshold membrane potential for generation of repetitive action potentials. However, a few assumptions had to be made for calculation, as not all data were available at that time. Later, Tatebayashi and Narahashi (1994) developed a method to calculate the percentage of sodium channel modification caused by pyrethroid based on patch clamp data using rat DRG neurons. Since the peak sodium current (INa) during a depolarizing pulse was not affected by pyrethroid, it represented the activity of normal or unmodified sodium channels. The tail current (Itail) upon termination of a depolarizing pulse appeared only after application of pyrethroid, and therefore it represented the activity of modified sodium channels. The percentage of modification (M) can be calculated by the following equation: M  [{I tail /( Eh  ENa )}/{I Na /( Et  ENa )}]  100

(1)

where Itail is the tail current amplitude obtained by extrapolation of the slowly decaying phase of the tail current to the moment of membrane repolarization assuming a single exponential decay, Eh is the potential to which the membrane was repolarized, ENa is the equilibrium potential for sodium ions obtained as the reversal potential for sodium current, and Et is the potential of step depolarization. The percentages of sodium channels modified by tetramethrin were very small: for example, for TTX-S sodium channels, 0.24%, 3.53%, and 12.03% by 0.1,1, and 10 M tetramethrin, respectively; for TTX-R sodium channels, 1.31%, 15.35%, 57.82%, and 81.20% by 0.01, 0.1, 1, and 10 M tetramethrin, respectively. Thus, TTX-R sodium channels are approximately 30 times more sensitive to tetramethrin than TTX-S sodium channels. A question arises as to the degree of pyrethroid modification needed to cause repetitive nerve activity. Using the same method of calculation and also comparing these calculated data with the threshold concentration for tetramethrin needed to induce repetitive discharges in rat cerebellar Purkinje neurons, an astonishingly small percentage was obtained, that is, 0.62%, as illustrated in Fig. 31.4 (Song and Narahashi, 1996). This provides one of the bases for high potency of pyrethroid action. It is also important to note that the significance of this ”toxicity amplification” is not limited to pyrethroids. When a drug slightly suppresses the slow depolarization (e.g., caused by activation of T-type calcium channels or in epileptic seizure), repetitive discharges generated by the slow depolarization will stop, and for this action only a concentration of the drug (e.g., antiepileptic drug) much lower than the IC50 for suppressing the depolarization (or calcium channels) will be needed, perhaps IC10 or even IC1. Thus, “pharmacological amplification” will become important for interpreting

Chapter | 31  Neurophysiological Effects of Insecticides

the drug action in vivo. The traditional concept of relating in vitro IC50 to a patient’s serum concentration of the drug may not necessarily be valid when the effect is exerted via the threshold phenomenon.

−110

0 TTX 0.5 µM Control

TM 0.3 µM TM 3 µM

2 nA 10 msec

TM 10 µM (a)

% Of Tetramethrin−Modified Channels

30 25 20 15 10 5 0 (b)

7

6

5 4 3 Tetramethrin (−log molar concentration)

Em (mV)

80 40 0 −40

50 msec

−80 (c) Figure 31.4  Concentration-dependent effect of tetramethrin on TTXS sodium currents of rat cerebellar Purkinje neurons, (a) Currents were evoked by a 5-msec step depolarization to 0 mV from a holding potential of 110 mV under control conditions and in the presence of tetramethrin (0.3 M, 3 M, and 10 M). TTX (0.5 M) completely blocked both the peak current and the tetramethrin-induced tail current, (b) The concentration-response relationship for induction of tail current. Each point indicates the mean  S.E.M.(n  6). Data were fitted by the Hill equation. The percentages of channels modified by tetramethrin are 0.62  0.15%, 2.19  0.36%, 5.75  0.87%, 13.58  1.35%, 22.77  2.26%, and 24.73  2.11% at concentrations of 0.1, 0.3, 1, 3, 10, and 30 M, respectively (n  6). (c) Repetitive after-discharges caused by 100 M tetramethrin, the threshold concentration. Action potentials were evoked by applying a current pulse (2 msec, 200 pA). Em refers to the membrane potential. From Song and Narahashi (1996).

803

Temperature Dependence of Pyrethroid Action It is well known that the insecticidal activity of pyrethroids and DDT increases with decrease in temperature. This is important, as the negative temperature dependence is partially responsible for selective toxicity in insects and mammals. This phenomenon is also deemed to contain some keys to the molecular mechanism of action of these insecticides. The earliest study was performed by Yamasaki and Ishii [Narahashi] (1954a) for the action of DDT on repetitive discharges of cockroach nerve. It was clearly demonstrated that the major factor for the negative temperature dependence of insecticidal action was the nerve sensitivity to DDT, which showed the Q10 value of 0.2. The effect of DDT in inducing repetitive discharges was reversible with respect to temperature change, and therefore, the metabolism of DDT did not come into play. Several studies have since been performed for the negative temperature dependence of DDT and pyrethroid actions on nerves (Ahn et al., 1987; Gammon, 1978; Narahashi, 1962; Salgado et al., 1989; Starkus and Narahashi, 1978). Binding of DDT to housefly brain increased with decreases in the temperature (Chang and Plapp, 1983). The sodium tail current slowed by pyrethroids was further slowed by lowering the temperature (Vijverberg et al., 1983). Despite these studies over many years, it was not until the mid-1990s that the physiological mechanism that underlies the negative temperature dependence of pyrethroid action on the nerve was clearly elucidated. Song and Narahashi (1996) have performed current clamp and voltage clamp experiments using rat cerebellar Purkinje neurons. Repetitive discharges induced by tetramethrin at 15–20°C subsided with an increase in the temperature to 30–35°C. The tail sodium channel current in the presence of tetramethrin was drastically affected by temperature changes (Fig. 31.5). Although the peak amplitude of the tail current was not changed by lowering the temperature from 30°C to 20°C, the decay phase of the tail current was greatly slowed, showing a Q10 value of 0.07, and the charge movement during tail current was increased, with a Q10 value of 0.2. Small Q10 values (large negative temperature dependence) for pyrethroid-induced tail current decay were also observed with frog nodes of Ranvier (Vijverberg et al., 1983). The percentage of sodium channels modified by tetramethrin was only slightly increased by lowering the temperature from 30°C to 20°C, with a Q10 value of 0.77. Thus, the most critical factor for the negative temperature dependence of repetitive discharges is slowing of the tail current decay, which causes a sizable increase in tail charge transfer by lowering the temperature. Selective Toxicity of Pyrethroids Pyrethroids are more toxic to insects than to mammals, with differences in LD50 ranging from 500- to 4500-fold (Elliott, 1977; Hirai, 1987; Miyamoto, 1993; Wiswesser, 1976). The selective toxicity of various insecticides has been generally ascribed to

Hayes’ Handbook of Pesticide Toxicology

804

35°C

Table 31.1  Factors Contributing Selective Toxicity of Pyrethroids

0 −110

Selectivity factor

Mammals

Insects

Differences

Due to temperature dependence

Low (37°C)

High (25°C) 5

Due to intrinsic sensitivity

Fast

High

10

Recovery

Fast

Slow

5

Due to enzymatic action

High

Low

3

Due to body size

High

Low

3

Potency on nerve 1 nA 10 msec

(a)

Detoxication rate

30°C Control

Overall difference  2250. (From Song and Narahashi, 1996.)

Tetramethrin 3 µM

(b)

25°C

(c)

20°C

*

(d) Figure 31.5  Temperature-dependent effect of 3 M tetramethrin on sodium currents recorded from a rat cerebellar Purkinje cell. The currents were evoked by a 5-msec step depolarization to 0 mV from a holding potential of 110 mV at various temperatures. The currents before and during application of tetramethrin are superimposed at each temperature. *, Current recording is truncated before the tail current returns to the baseline. From Song and Narahashi (1996).

higher rates of enzymatic degradation of insecticides in mammals than in insects. It was assumed that this was also the case for pyrethroids, albeit without any solid data to justify the assumption. However, this is not the case as far as pyrethroids are concerned. Various factors pertaining to selective toxicity of pyrethroids are given in Table 31.1. Pyrethroids are more potent on nerve function at low temperature than at high temperature, and the Q10 value is calculated to be 0.2, indicating that the potency increases fivefold with a decrease in temperature of 10°C, the body temperature difference between insects and mammals (Song and Narahashi, 1996). The intrinsic sensitivity of nerve (i.e., sodium channels) is at least 10 times, sometimes 100 times, higher in invertebrates than in mammals (Song and Narahashi, 1996; Warmke et al., 1997). Recovery after washout is approximately five times slower in invertebrates than in mammals (Song and Narahashi, 1996). Detoxication of pyrethroids is known to involve various enzymes whose rates are approximately three times lower in insects due to lower body temperature. Smaller body size in insects makes detoxication less efficient before pyrethroids reach the target site. A difference of 2250fold is obtained by multiplying differences in these factors, and this value is the same order of magnitude as the difference in LD50 described above. Therefore, the major factors responsible for the large difference in LD50 values between insects and mammals are all related to sodium channels. Vitamin E Alleviation of Pyrethroid-induced Pares­ thesia Any chemicals that block pyrethroid-modified sodium channels without effect on normal sodium channels

Neurophysiological Effects of Insecticides

31.1.2 Pyrethroid Action on Other Receptors and Channels Pyrethroid Modulation of GABAA Receptors Several papers have been published to report the block of GABAA receptors by type II pyrethroids (Abalis et al., 1986; Bloomquist and Soderlund, 1985; Crofton et al., 1987; Eshleman and Murray, 1990, 1991; Gammon and Sander, 1985; Lawrence and Casida, 1983; Lawrence et al., 1985; Lummis et al., 1987; Ramadan et al., 1988). Despite these reports, the matter has been controversial, as the potency and efficacy of pyrethroids in blocking the GABAA receptors are low. Our patch clamp experiments using rat DRG neurons have unequivocally shown that while 10 μM deltamethrin markedly prolongs the sodium current as expected, the GABA-induced current recorded from the very same neuron remains totally unchanged (Ogata et al., 1988). Thus, even though type II pyrethroids inhibit GABAA receptors to some extent under certain experimental conditions, the toxicological significance is rather questionable.

Cerebellar Purkinje TTX-S 25

CONTROL TM 10 μM +α − TOCO 30 μM TM 10 μM +α − TOCO 10 μM TM 10 μM 1 nA

(a)

% Of Modified Channels

could serve as antidotes for pyrethroid intoxication. Local anesthetics such as lidocaine were once considered (Oortgiesen et al., 1990), but they also block normal sodium channels. One such possibility is vitamin E, which has been used for prophylactic and therapeutic purposes to alleviate paresthesia caused by pyrethroids. The paresthesia includes tingling, itching, and burning sensation of the skin, without the clinical symptoms of erythema, edema, or vesiculation (Knox et al., 1984; LeQuesne et al., 1980). Pyrethroids, particularly type II pyrethroids, cause such paresthesia in the facial skin, and vitamin E has been used for therapeutic purposes (Flannigan and Tucker, 1985; Tucker et al., 1984). Vitamin E was found to be effective in blocking tetramethrin-modified sodium channels without effect on normal sodium channels in rat cerebellar Purkinje neurons and DRG neurons (Song and Narahashi, 1995). Vitamin E shortened the action potential duration prolonged by tetramethrin without affecting the peak amplitude. Reflecting this effect on action potential, the tail sodium channel current was blocked by vitamin E in a competitive manner while the peak sodium current remained unchanged (Fig. 31.6). However, the mechanism of this interesting antagonism is open to question, and it is not known whether the antioxidant action of vitamin E has anything to do with the antagonism. Recently, α-tocopherol was shown to antagonize the type I pyrethroid action in vivo in susceptible and kdr-resistant insects increases the LD50 values by 4.3 to 6.6-fold (Scott, 1998). Thus, α-tocopherol opens the door for development of antidotes for pyrethroid intoxication (Song and Narahashi, 1995).

805

TM 10 μM TM 10 μM +α − TOCO 10 μM TM 10 μM +α − TOCO 30 μM

20 15

(n=8)

10

10 msec

5 0

Concentration-Response Relationship 30

% Of Modified Channels

Chapter | 31

25 20 15 10 5 0

(b)

− α − TOCO (n=6) + α − TOCO (10 μM, n=6)

−7

−6

−5

−4

Tetramethrin (log molar concentration)

FIGURE 31.6 (a) Suppression of 10 μM tetramethrin-induced tail currents by 10 and 30 μM ()-a-tocopherol in TTX-S sodium channels of rat cerebellar Purkinje cells. Currents were evoked by depolarizing the membrane to 0 mV for 5 msec from a holding potential of 110 mV. Cells were first treated with 10 μM tetramethrin, and then 10 or 30 μM ()-α-tocopherol was added to the perfusion solution containing 10 μM tetramethrin. Records were taken 5 min after the addition of each chemical. The percentage of channel modification was calculated by Eq. (1). Mean  S.E.M. with n  6. (b) ()-α-Tocopherol shifts the concentration-response relationship for tetramethrin modification in the direction of higher concentrations in TTX-S sodium channels of cerebellar Purkinje cells. Mean  S.E.M.(n  6). From Song and Narahashi (1995).

Pyrethroid Modulation of Calcium Channels Permefhrin at a concentration as low as 50 pM increased the electrical activity of neurosecretory cells of the stick insect (Orchard and Osborne, 1979), and the effect was ascribed to the action on calcium channels (Gammon and Sander, 1985; Osborne, 1980). However, our patch clamp experiments using neuroblastoma cells showed a blocking, not a stimulating, action of pyrethroids on both T-type and L-type calcium channels (Yoshii et al., 1985). It should be noted that the observed impulse discharges from the insect neurosecretory cells may originate in sodium channels of presynaptic neurons.

806

Pyrethroid Modulation of Chloride Channels N1E-115 neuroblastoma cells are endowed with calcium-independent voltage-gated chloride channels. The type II pyrethroids deltamethrin and cypermethrin suppressed the channel activity by decreasing open probability, but type I pyrethroid cismethrin had much less effect (Forshaw et al., 1993; Ray et al., 1997). These chloride channels exhibited a high conductance of 340 pS. However, since the physiological function of these chloride channels is unknown, toxicological significance for pyrethroid action awaits further experimentation. Pyrethroid Modulation of Acetylcholine Receptors The binding of [3H]perhydrohistrionicotoxin to the Torpedo electric organ was inhibited by type I and type II pyrethroids (Abbassy et al., 1982, 1983a, b; Eldefrawi et al., 1984; Sherby et al., 1986). On the contrary, the frog endplate potential was not affected by allethrin (Wouters et al., 1977). This paradox remains to be solved. Pyrethroids have also been shown to interact with muscarinic ACh receptors (Eriksson and Nordberg, 1990; Eriksson and Fredricksson, 1991). Deltamethrin suppressed ACh-induced currents in Helix neurons (Kiss and Osipenko, 1991). A question was raised whether the action of pyrethroids on ACh receptors represented a specific interaction, because both active and inactive isomers of pyrethroids exerted nonspecific, inhibitory effects on the nicotinic ACh receptors of N1E115 neuroblastoma cells (Oortgiesen et al., 1989). The significance of ACh receptors, especially that of neuronal nicotinic ACh receptors, has received much attention these days with respect to physiology and pharmacology; thus, more elaborate experimental analyses for pyrethroid interactions with these receptors are warranted. Pyrethroid Modulation of Glutamate Receptors The [3H] kainate binding to mouse brain homogenates was inhibited by pyrethroids: IC50S were 80 nM for deltamethrin and 8 M for cispermethrin (Staatz et al., 1982). Cypermethrin at 1 M suppressed the glutamate sensitivity of the muscle of housefly larvae (Seabrook et al., 1988). However, the toxicological significance of glutamate receptors for pyrethroid action remains largely to be seen. Role of Calcineurin and Other Enzymes in Pyrethroid Action Pyrethroids have been shown to inhibit Na-Ca ATP hydrolysis and Ca-Mg ATP hydrolysis (Clark and Matsumura, 1987). Deltamethrin stimulated protein phosphorylation and caused the release of calcium from the intracellular storage sites (Enan and Matsumura, 1991; Matsumura et al., 1989). Pyrethroids, both type I and type II, stimulated phosphoinositide breakdown (Gusovsky et al., 1986). A striking discovery was made regarding calcineurin, neural calcium-calmodulin-dependent protein phosphatase, which was inhibited by type II pyrethroids such as cypermethrin, deltamethrin, and fenvalerate with IC50 values of 0.01–1 nM (Enan and Matsumura, 1992). By contrast, insecticidally inactive chiral isomers of these pyrethroids, active type I pyrethroids, DDT and heptachlor expoxide were much weaker inhibitors.

Hayes’ Handbook of Pesticide Toxicology

However, recent studies conducted by two independent groups cast doubt on the pyrethroid inhibition of calcineurin. None of the five pyrethroids tested, that is, bioallethrin, cyfluthrin, cypermethrin, deltamethrin, and fenvalerate, caused inhibition of the calcineurin-dependent dephosphorylation (Enz and Pombo-Villar, 1997). Both type I pyrethroids (cis-permethrin, trans-permethrin, and S-­bioallethrin) and type II pyrethroids (cis-cypermethrin, trans-cypermethrin, deltamethrin, and fenvalerate) were unable to inhibit the phosphatase activity of purified calcineurin (Fakata et al., 1998). Thus, the role of calcineurin in pyrethroid actions remains unclear.

31.1.3  Sodium Channel Mutation in Pyrethroid Resistance Earlier studies indicated that insecticide-resistant strains of insects acquired higher activity to detoxify insecticides (Wilkinson, 1983). However, a metabolic resistance mechanism could not completely explain the resistance to insecticides, because resistant strains of insects often contained unmetabolized insecticide in an amount much more than enough to kill susceptible strains. Insecticide resistance due to reduced nerve sensitivity was termed knockdown resistance (kdr) (Busvine, 1951; Milani, 1954). The mechanism of target site resistance was first studied for DDT, lindane and dieldrin. The sensitivity of the sensory nerves to DDT was lower in resistant houseflies than in susceptible houseflies (Smyth and Roys, 1955; Weiant, 1955). However, multiple discharges from the central nervous system (CNS) caused by insecticides are more closely related to the development of symptoms of poisoning. By electrophysiological measurements of such CNS multiple discharges as a measure of toxic action, resistant strains of houseflies were found to be less sensitive than susceptible strains for lindane and dieldrin (Yamasaki and Narahashi, 1958b) and for DDT (Yamasaki and Narahashi, 1962). While the identification of chromosome genes for low nerve sensitivity to DDT was made in the mid-1960s (Tsukamoto et al., 1965), studies for more precise sodium channel sites of mutations responsible for insecticide resistance were commenced only after thorough developments of molecular biology and genetic techniques in the 1990s. We now know mutations occur at several sites in the  subunit of sodium channels of pyrethroid-resistant kdr and super-kdr strains of various insects (Table 31.2).

31.2  Cyclodienes and hexachlorocyclohexane The mechanisms of action of cyclodienes and hexachlorocyclohexane (HCH) have a long history of studies. In the 1950s, dieldrin and lindane (-HCH) were shown to

Chapter | 31  Neurophysiological Effects of Insecticides

807

Table 31.2  Mutation of Sodium Channel Amino Acids in Pyrethroid Resistance Species

Locations of mutation in sodium channels

References

Domain and transmembrane segment

Amino acid sequence position

Drosophila melanogaster

IS4-S5

I253N

Pittendrigh et al. (1997)

Heliothis virescens

IS6

V410M

Park et al. (1997) Lee et al. (1999b)

Musca domestica

IIS4-S5

M918T

Williamson et al. (1996) Lee et al. (1999a)

Plutella xylostella

IIS5

T929I

Schuler et al. (1998)

Musca domestica

IIS6

L993F

Miyazaki et al. (1996) Williamson et al. (1996)

Blattella germanica

IIS6

L993F

Dong (1997) Miyazaki et al. (1996)

Heliothis virescens

IIS6

L993H

Park and Taylor (1997)

Musca domestica

IIS6

L1014F

Williamson et al. (1996) Smith et al. (1997) Lee et al. (1999a)

Blattella germanica

IIS6

L1014F

Miyazaki et al. (1996) Dong (1997)

Haematobia irritans

IIS6

L1014F

Guerrero et al. (1997)

Anopheles gambiae

IIS6

L1014F

Martinez-Torres et al. (1998)

Plutella xylostella

IIS6

L1014F

Schuler et al. (1998)

Heliothis virescens

IIS6

L1014H

Park and Taylor (1997)

Heliothis virescens

IIS6

L1029H

Lee et al. (1999b)

F, phenylalanine; H, histidine; I, isoleucine; L, leucine; M, methionine; N, asparagine; T, threonine; V, valine.

stimulate synaptic transmission in the cockroach nerve (Yamasaki and Ishii [Narahashi], 1954b; Yamasaki and Narahashi, 1958a). However, it was not until the 1980s that the GABA receptor was identified as their major target site by 36Cl uptake and [35S]t-butylbicyclophosphorothionate (TBPS) binding experiments (Abalis et al., 1986; Bermudez et al., 1991; Bloomquist and Soderlund, 1985; Bloomquist et al., 1986; Cole and Casida, 1986; Ghiasuddin and Matsumura, 1982; Llorens et al., 1990; Lummis et al., 1990; Matsumoto et al., 1988; Matsumura and Ghiasuddin, 1983; Pomes et al., 1994; Olsen et al., 1989; Thompson et al., 1990). The first electrophysiological experiment to demonstrate that the GABAA receptor was the target site was performed by Ogata et al. (1988), who showed lindane suppression of GABA-induced chloride currents in rat DRG neurons. The effects of lindane and dieldrin on single-channel

characteristics of cockroach GABA receptors were studied by noise analysis (Bermudez et al., 1991). Both insecticides decreased the frequency of channel opening. Dieldrin was without effect on the single-channel conductance, but lindane decreased it. However, Zufall et al. (1989) found no effect of lindane on single channels of crayfish stomach muscle. Whereas lindane inhibited all three types of GABAA receptors of rat cerebral cortex expressed in Xenopus oocytes, -, -, -HCH had differential effects (Woodward et al., 1992). Endrin, dieldrin, and lindane also suppressed electrophysiological responses of cockroach and locust GABA receptors (Bermudez et al., 1991; Wafford et al., 1989). Similarity between lindane and picrotoxin in blocking GABA receptors is pointed out (Tokutomi et al., 1994; Zufall et al., 1989). Dual Action of Dieldrin on GABAA Receptors Dieldrin has been found to exert a dual action on GABAA receptors.

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During repetitive co-applications of GABA and dieldrin, the GABA-induced current was first increased but later suppressed irreversibly (Fig. 31.7) (Nagata and Narahashi, 1994). The dual action was not only time dependent but also dieldrin concentration dependent. There were two components of suppression with IC50 values of 5 and 92 nM; EC50 for potentiation was 754 nM. Analysis of picrotoxin–dieldrin interaction experiments led to the conclusion that dieldrin acts on the picrotoxin site which is closely associated with the chloride channel. Dieldrin suppression of GABAA receptors accounts for its excitatory action, but the role of dieldrin potentiation remains to be seen.

31.2.1  GABAA Receptor Subunit Specificity of Dieldrin Action The GABAA receptor consists of five subunits which form a pentameric structure (Nayeem et al., 1994). There are at least six s, four s, four s including long and

Control

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3 min

4 min

5 min

500 pA 50 sec

Current amplitude (% of control)

(a)

(b)

250 200 150 100 50 0

0

5

10 15 Time (min)

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Figure 31.7  Effects of dieldrin on GABA-induced chloride currents in a rat dorsal root ganglion neuron, (a) Current records in response to 20sec application of 10 M GABA (solid bar) and to co-application of 10 M GABA and 1 M dieldrin (dotted bar) at the time indicated after taking control record. The peak amplitude of current was greatly enhanced but gradually decreased during repeated co-applications. Desensitization of current was accelerated, (b) Time course of the changes in peak current amplitude before and during (horizontal line) repeated co-applications. From Nagata and Narahashi (1994).

short splice variants of the 2 subunit, one , and one . Pharmacological sensitivity and profile are known to differ depending on the combination of these subunits. Dieldrin suppressed GABA-induced currents regardless of the three combinations, 12, 122S, and 622S, but dieldrin potentiation required the 2S subunit (Nagata et al., 1994). Dieldrin was more efficacious in potentiating the current in the 622S than in the 122S combination, indicating some role of the  subunits in potentiating the current. Whereas the 2 subunit was required for benzodi-azepine potentiation (Kurata et al., 1993; Pritchett et al., 1989) and zinc inhibition (Draguhn et al., 1990; Smart et al., 1991) of GABA responses, n-octanol potentiation did not require the 2 subunit (Kurata et al., 1993). Therefore, it is not possible to generalize the role of the 2 subunit in druginduced potentiation of GABA responses.

31.2.2  GabaA Receptor Subunit Specificity of The Actions of HCH Isomers HCH comprises geometric isomers which exhibit different insecticidal activity. The -HCH (lindane) is most toxic to mammals and insects and is a strong stimulant. The ­isomer is a weak stimulant, the -isomer is a weak depressant, and the -isomer is a strong depressant. Pomes et al. (1994) reported differential effects of HCH isomers on GABA-induced 36 Cl uptake by cortical neurons. Patch clamp experiments using rat DRG neurons showed differential actions of the four HCH isomers (Nagata and Narahashi, 1995). -HCH had a weak potentiating action and a strong inhibitory action on GABAinduced currents. -HCH had a strong potentiating action and an inhibitory action. -HCH and, -HCH had little or no effect on GABA-induced currents. The differential modulation of GABA response by HCH isomers accounts for variable symptoms of poisoning in insects and mammals. However, somewhat different results were obtained for the effects of HCH isomers on the 132S and 632S subunit combinations of GABAA receptors expressed in Xenopus oocytes (Aspinwall et al., 1997). GABA responses were inhibited by -HCH, potentiated by - and -HCH, and not affected by -HCH. Furthermore, the  sub-unit composition had no influence on these effects of HCH isomers. These differences in the responses to chemicals represent an example of the dissimilarity between native receptors and receptors expressed in Xenopus oocytes which is often encountered. -HCH altered calcium homeostasis and contractility of cardiac myocytes through interaction with ryanodine receptors (Buck and Pessah, 1999). -HCH also induced a profound increase in ionic permeability in lipid bilayers, and the calcium-dependent current produced by -HCH was selective for monovalent cations (K   Cs  Na) (Buck and Pessah, 1999).

Chapter | 31  Neurophysiological Effects of Insecticides

Cyclodiene Resistance The first direct demonstration of a target site resistance mechanism for dieldrin and lindane was reported by Yamasaki and Narahashi (1958b). Multiple discharges from the housefly CNS were induced by these insecticides, and resistant strains were less sensitive than susceptible strains. While low nerve sensitivity to dieldrin was also reported in resistant strains of Drosophila (Bloomquist et al., 1992; ffrench-Constant et al., 1991), it was not until 1993 that a point mutation in a Drosophila GABA receptor was found to be responsible for dieldrin resistance (ffrench-Constant et al., 1993). The cyclodiene resistance gene Rdl (resistance to dieldrin) was cloned from Drosophila resistant to cyclodienes and picrotoxinin. Single amino acid replacement from alanine to serine (A302S) occurs with the second membrane spanning domain, which is the region lining the chloride channel pore. Subsequently, similar mutations of amino acids were discovered in several other insect species resistant to dieldrin: in addition to A302S replacement in Drosophila melanogaster, A302G as well as A302S was found in Drosophila simulans, and a single mutation A302S also occurred in Aedes aegypti, Periplaneta americana, Musca domestica, and Tribolium castaneum (Anthony et al., 1998; Buckingham et al., 1996; Cole et al., 1995; ffrench-Constant, 1994; Miyazaki et al., 1995). In addition to Rdl, another GABA receptor subunit was also cloned from insects which represents a homolog of the vertebrate GABAA receptor  subunit. Contrary to the vertebrate GABAA receptor subunits, Rdl could form a functional homomultimeric receptor. The Rdl receptor was sensitive to the blocking action of picro-toxin but insensitive to that of bicuculline. GABA receptors formed by Rdl plus  subunits were insensitive to picrotoxin but sensitive to bicuculline (Zhang et al., 1995).

31.3  Fipronil Fipronil is a phenylpyrazole compound and was developed as a useful insecticide in the mid-1990s. It is effective against some insects such as the Colorado potato beetle and certain cotton pests that have become resistant to the existing insecticides. Fipronil is much more toxic to insects than to mammals, another advantage it has as an insecticide. Fipronil has been found to block insect GABA receptor (Rdl). Wild-type Rdl of Drosophila was suppressed by TBPS, 4-n-proply-4-ethynylbicycloorthobenzoate (EBOB), picrotoxinin, and fipronil (Buckingham et al., 1994a; Millar et al., 1994). Insect GABA receptors are different from vertebrate GABAA receptors in that they are not blocked by bicuculline (Benson, 1988; Buckingham et al., 1994a; ffrench-Constant et al., 1993; Millar et al., 1994; Sattelle et al., 1988), and are not potentiated by benzodiazepines and barbiturates (Millar et al., 1994). The insensitivity to

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bicuculline is reminiscent of the GABAC receptor of vertebrates (Qian and Dowling, 1993; Woodward et al., 1993). Dieldrin-resistant Drosophila melanogaster and D. simulans were also resistant to fipronil but to a much lesser extent, and the [3H]EBOB binding to these resistant strains was less inhibited by fipronil compared to susceptible strains (Cole et al., 1995). Mutant Drosophila Rdl (A302S) expressed in Xenopus oocytes was also less sensitive to fipronil than wild-type receptors (Hosie et al., 1995). Fipronil and desulfinyl derivative were more potent in houseflies than in mice as toxicants and in competing with [3H]EBOB binding (Hainzl and Casida, 1996). LD50 values of fipronil were 0.13 mg/kg and 41 mg/kg for housefly and mouse, respectively, and receptor IC50 values were 6.3 nM and 1010 nM for housefly and mouse, respectively. Fipronil block of GABAA receptors of rat DRG neurons has recently been analyzed in detail (Ikeda et al., 1999). Fipronil suppressed the GABA-induced wholecell currents reversibly with an IC50 of 1.66  0.18 M Preapplication of fipronil through the bath suppressed GABA-induced currents without channel activation. These results indicate that fipronil acts on the GABA receptors in the closed state. From co-application of fipronil and GABA, the IC50 value for the activated GABA receptor was estimated to be 1.12  0.21 M. The association rate and dissociation rate constants and the equilibrium dissociation constant of fipronil effect were estimated to be 673  220 M1 sec1, 0.018  0.0035 sec1, and 27 M for the resting GABA receptor, respectively, and 6600  380 M1 sec1, 0.11  0.0054 sec1, and 17 M for the activated GABA receptor, respectively. Thus, both the association and dissociation rate constants of fipronil for the activated GABA receptor are approximately ten times higher than those for the resting receptor, with a resultant lower Kd value for the activated receptor. Experiments with co-application of fipronil and picrotoxinin indicated that they did not compete for the same binding site. It is concluded that although fipronil binds to the GABAA receptor without activation, channel opening facilitates fipronil binding to and unbinding from the receptor. Single-channel recording experiments using the GABAA receptor of rat DRG neurons have revealed that fipronil prolonged the closed time without much effect on open time and burst du (Ikeda et al., 1999). Thus, fipronil reduces the frequency of channel opening, thereby suppressing the receptor activity.

31.4  Imidacloprid A number of factors must be taken into consideration for developing new insecticides and for using existing insecticides, mammalian toxicity and insecticide resistance being among the most important. In order to cope with the situation, a new group of chemicals has been developed into commercial

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insecticides during the past 10 or so years, that is, nitromethylene or chloronicotinyl insecticides. Soloway et al. (1978, 1979) found that nithiazin was among the most insecticidally active nitromethylenes tested. Imidacloprid was later shown to have excellent insecticidal activities against leafhoppers, planthoppers, white flies, aphids, and various coleopteran insects (Elbert et al., 1990, 1991). Newer derivatives of imidacloprid were also developed, including nitenpyram (Minamida et al., 1993a, b) and acetamiprid (Takahashi et al., 1992). Imidacloprid exhibits a unique mechanism of action on nicotinic acetylcholine (nACh) receptors. It bound to insect nACh receptors with a high affinity (Bai et al., 1991; Buckingham et al., 1995; Chao and Casida, 1997; Chao et al., 1997; Lind et al., 1998; Liu and Casida, 1993; Liu et al., 1994). Imidacloprid depolarized nerve membrane and caused spontaneous discharges in cockroaches (Buckingham et al., 1995; Nishimura et al., 1994, 1998; Sone et al., 1994). Imidacloprid, acetamiprid, and nitenpyram also acted on Torpedo nACh receptors, but only as weak agonists (Tomizawa et al., 1995). Mammalian endplate nACh receptors were less sensitive than those of locust neurons (Zwart et al., 1994). The effects of imidacloprid on single-channel activity of nACh receptors were analyzed in detail using PC12 cells (Nagata et al., 1996, 1997, 1998). First, whole-cell currents were analyzed in the absence and presence of imidacloprid. Imidacloprid itself generated whole-cell currents with a low potency and efficacy. The minimum ­ effective concentration was 1 M, and the current amplitude reached a maximum at 30 M. The imidacloprid-induced current was approximately 10% of the carbachol-induced current. Imidacloprid also suppressed carbachol-induced currents with a low potency: even at the maximum concentration tested (100 M), imidacloprid suppressed the currents only by 30%. Single-channel analyses have disclosed an interesting feature of imidacloprid action. Application of ACh induced primarily main conductance (25.4 pS) currents and some low conductance (9.8 pS) currents, while imidacloprid generated primarily the low conductance currents (Fig. 31.8a and b). Co-application of ACh and imidacloprid generated both types of currents (Fig. 31.8c). The mean open time and burst duration of the main conductance current were decreased by the co-application of ACh and imidacloprid. These changes in single-channel behavior by imidacloprid can account for the changes in whole-cell ACh receptor currents. Imidacloprid has both agonist and antagonist effects on the mammalian neuronal nicotinic ACh receptors. Nitenpyram behaved similarly to imidacloprid in modulating single ACh-induced currents of PC12 cells (Nagata et al., 1999). ACh receptor subunit specificity for imidacloprid action has recently been studied (Matsuda et al., 1998). Imidacloprid was a partial agonist in generating currents in the recombinant chicken 42 subunit combination and in the hybrid receptor of Drosophila  subunit (SAD)

50 ms

10 µM ACh

5 pA

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Sub open

(a) 10 µM Imidacloprid

Close

50 ms 5 pA

Sub open

Main open

(b) 10 µM ACh + 10 µM Imidacloprid

50 ms 5 pA

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(c) Figure 31.8  Single-channel currents activated by 10 M ACh, 10 M imidacloprid, and co-application of 10 M ACh and 10 M imidacloprid to cell-attached membrane patches clamped at a membrane potential 40 mV more positive than the resting potential in PC12 cells, (a) Currents induced by 10 M ACh occurred during brief isolated openings or longer openings interrupted by a few short closures or gaps. Main conductance state currents were observed more frequently than subconductance state currents. (b) Currents induced by 10 M imidacloprid. Subconductance state currents were more frequently observed than main conductance state currents. (c) Co-application of 10 M ACh and 10 M imidacloprid. Main conductance and subconductance state currents were induced, and channel openings were shortened. From Nagata et al. (1998).

with the chicken 2 subunit, both expressed in Xenopus oocytes. However, imidacloprid was more potent on the SAD2 subunit combination than on the 42 combination. Furthermore, imidacloprid was a weak potentiator of ACh-induced currents in the 42 receptors, whereas it was a weak antagonist of ACh-induced currents in the SAD2 receptors. Binding experiments indicated that imidacloprid, acetamiprid, and nitenpyram had low to moderate potency at the 3 and 42 ACh receptors and were essentially inactive at the 1 and 7 ACh receptors (Tomizawa and Casida, 1999). Insect ACh receptor subunits were also studied for imidacloprid action (Huang et al., 1999). In the peach-potato aphid Myzus persicae, five  subunit cDNAs have been cloned: Mp1, Mp2, Mp3, Mp4, and Mp5. Although the insect  subunits evolved in parallel with the vertebrate

Chapter | 31  Neurophysiological Effects of Insecticides

neuronal nACh receptors, the insect non- subunits are different from vertebrate neuronal  and muscle non- subunits. The aphid nACh receptor  sub-unit cDNAs were co-expressed with the rat 2 subunit in Drosophila S2 cells. The affinity of recombinant nACh receptors for [3H] imidacloprid was a subtype dependent, being high in Mp2 and Mp3 subunits, but low in Mpa1 subunit.

Acknowledgments Author’s studies quoted in this chapter were supported by NIH Grant NS14143. Thanks are also due to Julia Irizarry for secretarial assistance and to Nayla Hasan for technical assistance.

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Strichartz, G. (1973). The inhibition of sodium currents in myelinated nerve by quaternary derivatives of lidocaine. J. Gen. Physiol. 62, 37–57. Tabarean, I. V., and Narahashi, T. (1998). Potent modulation of tetrodotoxin-sensitive and tetrodotoxin-resistant sodium channels by the type II pyrethroid deltamethrin. J. Pharmacol. Exp. Ther. 24, 958–965. Takahashi, H., Mitsui, J., Takakusa, N., Matsuda, M., Yoneda, H., Suzuki, J., Ishimitsu, K., and Kishimoto, T. (1992). NI-25, a new type of systemic and broad spectrum insecticide. Brighton Crop Protection Conferences—Pest and Diseases 1, 89–96. Takeda, K., and Narahashi, T. (1988). Chemical modification of sodium channel inactivation: separate sites for the action of grayanotoxin and tetramethrin. Brain Res. 448, 308–312. Tanguy, J., and Yeh, J. Z. (1991). BTX modification of Na channels in squid axons. I. State dependence of BTX action. J. Gen. Physiol. 97, 499–519. Tatebayashi, H., and Narahashi, T. (1994). Differential mechanism of action of the pyrethroid tetramethrin on tetrodotoxin-sensitive and tetrodotoxinresistant sodium channels. J. Pharmacol. Exp. Ther. 270, 595–603. Thompson, R. G., Menking, D. E., and Valdes, J. J. (1990). Comparison of lindane, bicyclophosphate and picrotoxin binding to the putative chloride channel sites in rat brain and Torpedo electric organ. Neurotoxicol. Teratol. 12, 57–63. Tokutomi, N., Ozoe, Y., Katayama, N., and Akaike, N. (1994). Effects of lindane (-BHC) and related convulsants on GABAA receptor-operated chloride channels in frog dorsal root ganglion neurons. Brain Res. 643, 66–73. Tomizawa, M., and Casida, J. E. (1999). Mino structural changes in nicotinoid insecticides confer differential subtype selectivity for mammalian nicotinic acetylcholine receptors. Brit. J. Pharmacol. 127, 115–122. Tomizawa, M., Otsuka, H., Miyamoto, X., and Yamamoto, I. (1995). Pharmacological effects of imidacloprid and its related compounds on the nicotinic acetylcholine receptor with its ion channel from the Torpedo electric organ. J. Pesticide Sci. 20, 49–56. Trainer, V. L., McPhee, J. C, Boutelet-Bochan, H., Baker, C., Scheuer, T., Babin, D., Demoute, J.-P, Guedin, D., and Catterall, W. A. (1997). High affinity binding of pyrethroids to the  subunit of brain sodium channels. Mol. Pharmacol. 51, 651–657. Tsukamoto, M., Narahashi, T., and Yamasaki, T. (1965). Genetic control of low nerve sensitivity to DDT in insecticide-resistant houseflies. Botyu-Kagaku (Scientific Pest Control) 30, 128–132. Tucker, S. B., Flannigan, S. A., and Ross, C. E. (1984). Inhibition of cutaneous paresthesia resulting from synthetic pyrethroid exposure. Internal. J. Dermatol. 23, 686–689. Vijverberg, H. P. M., and van den Bercken, J. (1990). Neurotoxicological effects and the mode of action of pyrethroid insecticides. Critical Reviews in Toxicology 21(2), 105–126. Vijverberg, H. P. M., van der Zalm, J. M., and van den Bercken, J. (1982). Similar mode of action of pyrethroids and DDT on sodium channel gating in myelinated nerves. Nature 295, 601–603. Vijverberg, H. P. M., van der Zalm, J. M., van Kleef, R. G. D. M., and van den Bercken, J. (1983). Temperature- and structure-dependent interaction of pyrethroids with the sodium channels in frog node of Ranvier. Biochim. Biophys. Acta 728, 73–82. Wafford, K. A., Sattelle, D. B., Gant, D. B., Eldefrawi, A. T, and Eldefrawi, M. E. (1989). Noncompetitive inhibition of GABA receptors to insect and vertebrate CNS by endrin and lindane. Pesticide Biochem. Physiol. 33, 213–219. Warmke, J. W., Reenan, R. A., Wang, P., Qian, S., Arena, J. P., Wang, J., Wunderler, D., Liu, K., Kaczorowski, G. J., Van der Ploeg, L. H.,

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Chapter | 31  Neurophysiological Effects of Insecticides

Yeh, J. Z. (1979). Dynamics of 9-aminoacridine block of sodium channels in squid axons. J. Gen. Physiol. 73, 1–21. Yeh, J. Z. (1980). Blockage of sodium channels by stereoisomers of local anesthetics. In “Molecular Mechanisms of Anesthesia,” (B. R. Fink ed.), pp. 35–44. Raven Press, New York. Yeh, J. Z. and Narahashi, T. (1977). Kinetic analysis of pancuronium interaction with sodium channels in squid axon membranes. J. Gen. Physiol. 69, 293–323. Yoshii, M., Tsunoo, A., and Narahashi, T. (1985). Effects of pyrethroids and ve-ratridine on two types of Ca channels in neuroblastoma cells. Soc. Neurosci. Abstr. 11, 518.

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Chapter 32

The Influence of Age on Pesticide Toxicity Carey Pope Department of Physiological Sciences, Center for Veterinary Health Sciences, Oklahoma State University, Stillwater, Oklahoma

32.1  General concepts in differential sensitivity to pesticides Age-related differences in sensitivity to pesticides can be based on a number of factors. Toxicokinetic differences among age groups can contribute to differential sensitivity, with differences in biotransformation often being a major factor. In other instances, toxicodynamic differences may exist which lead to age-related differences in sensitivity. For example, during development and maturation, a critical time of exposure or “window of opportunity” during which a developmental process occurs may impart selective sensitivity. At the other end of the spectrum, changes associated with aging may alter sensitivity to pesticides. Moreover, the relative contribution of toxicokinetic and toxicodynamic factors in age-related sensitivity may differ markedly among the various classes of pesticides, and even among members of the same class of toxicants. In contrast to toxicokinetic and toxicodynamic differences, exposures to pesticides can often be markedly different among age groups, based on age-specific behaviors, diets, or other factors. Thus, the nature of age-related differences in sensitivity to pesticides is complex, and broad-based generalities are typically unjustified, even within the same class of agents (Table 32.1). With the common routes of exposure (i.e., oral, dermal, and inhalation), a pesticide must first be absorbed before systemic toxicity can occur. Knaak and coworkers (1984) reported a doubling of dermal absorption rate for triadimefon in young male rats compared to adults. In a comparative study of 14 different pesticides, 11 of these exhibited agerelated differences in percutaneous absorption (Shah et al., 1987). Interestingly, 4 of the 14 showed greater absorption in young (33-day-old) while 7 of the 14 showed more extensive absorption in adult (82-day-old) animals. Moreover, even within the same class of pesticide (e.g., the organophosphorus toxicants parathion and chlorpyrifos), no clear age-related pattern of dermal absorption was evident, Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

Table 32.1  General Factors Contributing to Age-related Differences in Sensitivity to Pesticides Toxicokinetic

Differences in absorption Differences in distribution/elimination Differences in biotransformation

Toxicodynamic

Age-related expression of target molecules or sensitive processes Differential capacities to recover from or adapt to toxicant insult

Exposure-based

Age-related behaviors Age-related diets Differences in time available for cumulative exposures and/or expression of toxicity

i.e., chlorpyrifos showed greater absorption in young while parathion showed greater absorption in older animals. Hall and coworkers (1992) reported that dermal absorption of the dinitrophenol pesticide dinoseb was lower (about 20%) in 33-day-old rats compared to adults (82 days of age). Very similar rates of dermal absorption in young and adult female rats were reported for the organochlorine pesticide chlordecone (Heatherington et al., 1998). It should be noted, however, that a number of these studies used the same age groups, i.e., 33 and 82 days of age, to represent young and adult rats, respectively. Thus, differences in absorption for even younger animals are relatively unknown. Rate or extent of absorption can likely contribute, however, to differential sensitivity among age groups in some cases. Once absorbed, differences in tissue distribution or rates of elimination between age groups can contribute to differential sensitivity. Concentrations of the fungicide captan in kidney and liver were lower in young rats than adults following equivalent absorbed dosages (Fisher et al., 1992). Older animals (and people) typically have higher fat

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content than younger individuals, which can have an important effect on distribution, accumulation and storage of highly lipophilic pesticides, e.g., organochlorines. Deichmann (1972) reported that DDT was eliminated from the body most efficiently in neonates and less so in older rats, at least partially because of differences in partitioning of the pesticide into fatty tissues. Obviously, differences in fat content can affect not only short-term distribution into tissues, but also long-term distribution and deposition of nonpolar pesticides. For example, the concentrations of DDT, DDE and several PCB congeners were reported to increase in an age-related manner in human tissues at autopsy (Park et al., 2005). Changes in biotransformation during maturation and aging can be critically important in age-related differences in sensitivity to pesticides. Immature and very old animals generally have lower biotransformation capacities, e.g., lower levels of cytochrome P450-dependent monooxygenases (Benke and Murphy, 1975; Mehendale, 1980; Wynne et al., 1987). If a pesticide is activated by cytochrome P450-dependent monooxygenases to a more toxic metabolite, lower levels of cytochrome P450-dependent monooxygenases could potentially be associated with lower sensitivity to that pesticide. By contrast, pesticides which are effectively inactivated by monooxygenases could be relatively more toxic in groups with lower levels of cytochrome P450-dependent monooxygenases. Lower activities of phase II reactions in neonatal or aged animals may also increase sensitivity to certain pesticides (Borghoff et al., 1988; Das et al., 1981; Egaas et al., 1995; Fujita et al., 1985; Jang et al., 2001). Because of the complexity of pathways and the multiplicity of reactions generally involved in xenobiotic metabolism, however, differences in individual metabolic processes between age groups have to be considered in context to appreciate the net consequences of biotransformation on age-related toxicity. For example, while young rats exhibit lower rates of cytochrome P450dependent monooxygenases-mediated activation of the organophosphorus pesticide parathion to its active metabolite, paraoxon, lesser capacity in neonates for detoxification of paraoxon appears to be a prominent difference contributing to higher sensitivity in younger rats (Atterberry et al., 1997; Benke and Murphy, 1975; Karanth and Pope, 2000). Toxicodynamic differences can also contribute to agerelated sensitivity. The ability to restore function following toxicant exposure may be higher in some age groups than in others. For example, young rats challenged with hepatotoxicants recover much better than older animals, apparently because of more rapid and robust synthesis of new cells following the initial tissue damage (Dalu and Mehendale, 1996). More rapid recovery of acetylcholinesterase (AChE) activity in younger animals (Chakraborti et al., 1993; Moser, 1999; Pope and Liu, 1997; Pope et al., 1991) and slower recovery in aged animals (Michalek et al., 1990) following acute exposure to an organophosphorus

Hayes’ Handbook of Pesticide Toxicology

anticholinesterase may make those age groups differentially sensitive to accumulative inhibition with subsequent exposures to the same or different anticholinesterases. Age-related differences in sensitivity to pesticides can be influenced by multiple toxicokinetic and toxicodynamic factors. An important consideration in the differential sensitivity to pesticides can be the time available for toxicity to develop. Children have a longer time to live than adults, thus if pesticide exposure leads to the development of some form of delayed toxicity, e.g., tumor formation, a child has more time for this adverse effect to be exhibited. Conversely, older individuals have experienced a longer time to accumulate residues of persistent pesticides or damage from chronic exposures. As noted before, the critical time-dependent nature of developmental stages is also an important consideration in age-related differences in response to pesticides. The endogenous metabolite bilirubin, for example, induces encephalopathy in the developing nervous system only at certain early timepoints when the blood–brain barrier is deficient (Lee et al., 1995; Wennberg, 1993). Another factor of particular importance to age-related differences in susceptibility is differential exposures. Agerelated behaviors may contribute to differential exposure and sensitivity. For example, young children tend to sample the environment by taste. If the opportunity arises for oral “sampling” of a pesticide container, the young child may be much more susceptible to toxicity based on a greater likelihood of such exposure. In general, young children tend to be more exploratory and inquisitive than adults, which can sometimes lead to contact with inappropriately stored chemicals. Many lipophilic xenobiotics concentrate in breast milk, thus breast-feeding infants may be preferentially exposed to such toxicants (Mussalo-Rauhamaa et al., 1984; Schildkraut et al., 1999). Young children eat more in proportion to their body size and they tend to eat more frequently than adults. When pesticide residues are consumed with the food, the relative frequency of exposure can be important if recovery takes longer than the time between exposures. Toddlers are also in contact with the floor more than adults. With a higher surface area:body weight ratio, dermal contact may be more extensive than in adults. When pesticide residues fall to the floor after household applications or become associated with carpeting or furniture, there is a higher probability of direct dermal contact in children playing on those surfaces (Fenske et al., 1990; Lu and Fenske, 1999). Conversely, adults can be exposed to chemicals in the workplace, an exposure possibility which is generally missing in young children and older adults. Obviously, there are many reasons why exposures to pesticides can be age-related. The role of differential exposure in age-related sensitivity to pesticides is a critical issue and is discussed in more detail in later chapters. It is apparent, however, that age-related differences in sensitivity to pesticides can be caused by either differences in

Chapter | 32  The Influence of Age on Pesticide Toxicity

inherent sensitivity to the pesticide, differences in exposure, or both. Clearly, multiple factors can contribute to differential susceptibility to pesticides throughout life. Risk assessment for pesticides relies heavily on data generated from animal studies. The United States Environmental Protection Agency previously prohibited the use of human data in the pesticide registration process (U.S. EPA, 1998a). In 2006, however, U.S. EPA published a ruling for protecting human subjects in such research. A Human Studies Review Board was subsequently established to evaluate scientific and ethical aspects of research proposals and reports of completed research with human data. Thus, with appropriate review, human data can now be part of the registration process. Obviously, animal models continue to provide the primary experimental information supporting pesticide registration. The use of rodent animal models to estimate age-related differences in sensitivity in humans has some inherent problems however, in particular when modeling the effects of early postnatal exposures. Developmentally, the maturational states of experimental animals and humans at parturition and perinatal periods can be quite different (Romijn et al., 1991). If neonatal rodents are more sensitive than adults to a particular pesticide, but only briefly during the early postnatal period, they may not represent a valid model for children because of species differences in maturation relative to the timing of exposure. The comparative development, maturation, and aging of organ systems between man and experimental animals must be kept in mind when extrapolating age-related differences in sensitivity from animal models.

32.2  Children’s health and regulation of pesticides in the United States Ideally, regulatory policies governing the use of pesticides should be conservative enough to allow for protection of all members of the population. With noncarcinogenic toxicants, an uncertainty factor of ten has been traditionally incorporated into the risk assessment process for such purposes (Barnes and Dourson, 1988), assuming that variability in sensitivity to a particular agent within subpopulations is no greater than an order of magnitude. For this to be true and for all members of the population to be protected, all possible extrinsic and intrinsic modifiers of toxicity, e.g., nutrition, disease, physiological stressors, genetic polymorphisms, etc., must together contribute to less than a 10-fold variation in sensitivity in the entire population. One intrinsic modifier of pesticide toxicity, age, has received considerable attention in recent years. In particular, the relative sensitivity of developing infants and children to pesticides has been the focus of concern (Bearer, 1995; Bellinger, 2007; Fenner-Crisp, 1995; Garrettson, 1997; Goldman, 1995; Jurewicz et al., 2006; Little, 1995; Rosas

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and Eskenazi, 2008; Tilson, 1998). In 1988, the National Academy of Sciences (NAS) initiated a concerted effort to evaluate pesticide exposures in infants and children and to determine if the health of children was adequately addressed in the regulation of pesticides. The Committee on Pesticides in the Diets of Infants and Children, composed of scientists from industry, government and academia, was established within the National Research Council of NAS in 1988 to evaluate the relative sensitivity of infants and children to pesticides. The conclusions eventually reached by this select committee had far-reaching consequences (see later). In 1989, public attention in the United States was focused on the possibility that children were being exposed to excessive levels of pesticide residues in food products by media coverage of a report from the Natural Resources Defense Council (NRDC) entitled “Intolerable Risk: Pesticides in Our Children’s Food” (NRDC, 1989). The executive summary of this report begins “Our nation’s children are being harmed by the very fruits and vegetables we tell them will make them grow up healthy and strong.” While the basis of many claims in the NRDC report may have been inaccurate (Wilkinson and Ginevan, 1989), the public attention raised by this report had a significant impact, i.e., it strengthened the commitment to ensure that children’s health was adequately considered in the risk assessment of pesticides. Four years later, the National Academy of Sciences published the report “Pesticides in the Diets of Infants and Children” (NAS, 1993) which detailed conclusions from the NRC committee with the same name. Major findings of this committee included: (1) both quantitative and qualitative differences in toxicity of pesticides can occur between children and adults but quantitative differences are usually less than a factor of ten; (2) infants and adults differ quantitatively and qualitatively in the types of pesticide exposures in the diet, a factor of generally more importance than differences in inherent sensitivity; (3) that assessment of pesticide exposures should consider dietary as well as nondietary sources; and (4) that “in the absence of data to the contrary, there should be a presumption of greater toxicity to infants and children” (NAS, 1993). The findings from this committee provided impetus for federal legislation addressing pesticide regulation, in particular regarding potential problems with differential exposure and sensitivity in children. In 1996, the Food Quality Protection Act (FQPA) was passed into law containing sections relating to the protection of infants and children from pesticide exposures. The FQPA amended the Federal Insecticide, Fungicide and Rodenticide Act and the Federal Food, Drug and Cosmetic Act (FFDCA). Section 408(b)(2)(C) of FFDCA states that with “threshold” adverse effects, “an additional tenfold margin of safety for the pesticide chemical residue … shall be applied for infants and children to take into account

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potential pre- and post-natal toxicity and completeness of the data with respect to exposure and toxicity to infants and children.” Further, this section of FFDCA states that the “Administrator may use a different margin of safety for the pesticide chemical residue only if, on the basis of reliable data, such margin will be safe for infants and children.” In October of 1995 the U.S. EPA announced that it would explicitly evaluate risks to infants and children in all regulatory actions and in April of 1997, Executive Order 13045 directed Federal agencies to identify and assess environmental health and safety risks to children (U.S. EPA, 1998b). Thus, the default position of the U.S. EPA in pesticide regulatory decisions was to use an additional 10 uncertainty factor (the FQPA factor) for threshold effects to insure the protection of infants and children from pesticide toxicity. The U.S. EPA Office of Pesticide Programs proposal included, however, the possibility of either removing or reducing the magnitude of the FQPA factor if “reliable data” were available that suggested infants and children would be adequately protected under those conditions (U.S. EPA, 1999). Thus, risk assessment procedures for pesticides registered with the U.S. EPA now incorporate an additional FQPA uncertainty factor for infants and children unless sufficient data indicate that the young are not at higher risk. The conclusions from the NAS report (NAS, 1993) regarding risks to infants and children were based on two parameters, i.e., differences in sensitivity and differences in exposure. In practice, the decision to use a 10-fold FQPA safety factor or something else is driven by the risk characterization process and is not determined in the derivation of the reference dose. A “weight-of-evidence” approach is taken based on all hazard and exposure data, considering the level of confidence in these assessments and thus any residual uncertainties. The following is a brief summary of evidence pertaining to age-related differences in response to pesticides. It should be noted that while the recent focus of concern in the United States has been on the possibly higher susceptibility of infants and children, because of the demographics of societal aging, elderly individuals and their relative susceptibility to pesticides could become a more important issue (Overstreet, 2000). Alterations in cholinergic neurotransmission with aging and associated neurological disorders such as Alzheimer’s disease may be particularly important in contributing to differential sensitivity to the cholinesterase-inhibiting agents and with other pesticides which may alter cholinergic neurotransmission.

32.3  Age-related differences in sensitivity to pesticides It is apparent that, as with other types of xenobiotics (Done, 1964; Goldenthal, 1971), there is no consistent

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effect of age on acute sensitivity to pesticides across all chemical classes or even within a class of compounds. There are various factors that could contribute to differential toxicity, whether one compares different age groups, different species, different sexes, different strains, or with any other comparison. These contributing factors will be examined in more detail with specific examples of pesticides potentially capable of eliciting age-related effects (Table 32.2).

32.3.1  Organophophorus Pesticides Organophosphorus pesticides (OPs) elicit toxicity through inhibition of AChE (Mileson et al., 1998). Age-related differences in sensitivity to OPs have been reported in many experimental studies (Benke and Murphy, 1975; Brodeur and DuBois, 1963; Gagne and Brodeur, 1972; Gaines and Linder, 1986; Mendoza, 1976; Moser and Padilla, 1998; Pope et al., 1991). In general (but not always), neonatal animals are more sensitive to the acute toxicity of OPs. Lu and coworkers (1965) reported a maturational decrease in sensitivity to malathion among newborn, 14- to 16-day-old and adult rats. Mendoza (1976) reported that 1-day-old rats were about nine times more sensitive to lethality from acute exposure to malathion. Mortality in newborn pigs following dermal application of chlorpyrifos (2.5% aerosol) was markedly higher when exposure occurred within the first 3 h of life than at 30–36 hours after birth, suggesting a rapid change in sensitivity in the first days following parturition (Long et al., 1986). Pope and coworkers (1991) reported that 7-day-old rats were between two and nine times more sensitive than adult (90 days of age) rats to the acute toxicity of methyl parathion, parathion, and chlorpyrifos. Diazinon also appears more acutely toxic in young rats compared to adults (Padilla et al., 2004). By contrast, methamidophos appears to elicit little age-related toxicity (Moser, 1999; Padilla et al., 2000). Several factors could contribute to age-related differences in response to acute OP exposures. Gagne and Brodeur (1972) investigated potential metabolic factors in the higher sensitivity of weanling rats to parathion and concluded that limited detoxification of parathion and its metabolite paraoxon were at least partially responsible. Later, Benke and Murphy (1975) evaluated metabolic contributions to age-related differences in sensitivity to parathion and methyl parathion. When biotransformation of parent and metabolites of parathion and methyl parathion was compared to LD50 values among different age groups, high correlations were noted between lethality and liver and plasma A-esterase activity, oxon dealkylation and dearylation, and binding to “noncritical tissue constituents” in liver and plasma. They concluded that more robust metabolic inactivation of the active oxons of these two pesticides in more mature animals was primarily responsible for

Chapter | 32  The Influence of Age on Pesticide Toxicity

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Table 32.2  Studies Reporting Age-related Differences in Sensitivity with the Major Classes of Pesticides Pesticide class

Type of study: Relative sensitivity

Citations

Organophosphorus

Animal study: Immature more sensitive than adults

Atterberry et al. (1997); Benke and Murphy (1975); Brodeur and DuBois (1963); Gagne and Brodeur (1972); Karanth and Pope (2000); Long et al. (1986); Mendoza (1976); Moser and Padilla (1998); Padilla et al. (2004); Pope et al. (1991)

Organophosphorus

Animal study: Adults more sensitive than immature

Chakraborti et al. (1993); Harp et al. (1997); Johnson and Barnes (1970); Lu et al. (1965); Moretto et al. (1991); Peraica et al. (1993); Pope and Liu (1997); Pope et al. (1992, 1993)

Organophosphorus

Animal study: Aged adults more sensitive than young adults

Karanth and Pope (2000); Karanth et al. (2006); Veronesi et al. (1990)

Organophosphorus

Human study: Children more sensitive than adults

Diggory et al. (1977)

Organochlorines

Animal study: Immature more sensitive than adults

Eriksson (1997); Jinna et al. (1989); Samanta and Chainy (1997)

Organochlorines

Animal study: Adults more sensitive than immature

Kiran and Varma (1988); Lu et al. (1965)

Carbamates

Animal study: Immature more sensitive than adults

Moser (1999) (based on lethality)

Carbamates

Animal study: Aged adults more sensitive than young adults

Knisely and Hamm (1989); Takahashi et al. (1991)

Carbamates

Human study: Children more sensitive than adults

Lifshitz et al. (1997) (depending on endpoint)

Pyrethroids

Animal study: Immature more sensitive than adults

Cantalamessa (1993); Sheets et al. (1994)

the relative decrease in sensitivity with age. Atterberry and coworkers (1997) compared the toxicity and biotransformation of parathion and chlorpyrifos in neonatal and adult rats and concluded that differences in liver carboxylesterase activity and cytochrome P450-dependent monooxygenasesdependent dearylation were important in differential agerelated sensitivity to these pesticides. Moser and colleagues (1998) concluded that differences in liver carboxylesterase and A-esterase activities formed the basis for age-related differences in sensitivity to acute chlorpyrifos exposures. Other studies have indicated that maturational differences in the capacity for detoxification of organophosphates by A-esterases and carboxylesterases may contribute to higher sensitivity to these pesticides in immature animals (Costa et al., 1990; Li et al., 1993, 1995; Maxwell, 1992; Pond et al., 1995). Padilla and coworkers (2004) proposed that an in vitro screen that evaluated the age-related carboxylesterase and A-esterase detoxifying activities for a particular OP could be used to predict in vivo sensitivity. Karanth and Pope (2000) reported that plasma carboxylesterase correlated highly with acute sensitivity to parathion in neonatal, juvenile, adult and aged rats. Thus, considerable evidence suggests that immature animals are more sensitive to the acute toxicity of several OP pesticides

because of limited detoxification of either the parent compound or its active metabolite, and that this may contribute to differential sensitivity in aging. Young children also appear to be more sensitive to acute toxicity from OP exposure. In a case of parathioncontaminated food in Jamaica, the highest incidence of lethality was in children less than 5 years of age (Diggory et al., 1977). Differences in metabolic capacities between very young children and older children or adults may also be primary determinants in age-related sensitivity to acute OP exposures. Augustinsson and Barr (1963) showed that serum arylesterase (A-esterase) activity was very low in newborn children but increased steadily during the first 6 months of life. Ecobichon and Stephens (1973) reported that plasma cholinesterase and A-esterase activities increased dramatically in children during the first year of life, after which no further increases occurred. Any active anticholinesterases in the blood of very young children would therefore be less likely to bind to nontarget cholinesterases or to be hydrolyzed by A-esterases, thus more inhibitor would be available to reach target tissues. As detoxification of active OP anticholinesterases is thought to be a prominent factor in age-related sensitivity (Atterberry et al., 1997; Benke and Murphy, 1975;

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Mortensen et al., 1996; Moser et al., 1998), these studies suggest that dramatically higher acute sensitivity in children may only exist in the very young (1 year of age), however, when these detoxification processes appear most limited. A more recent study suggested that total liver carboxylesterase activity in human infants may in fact change relatively little after the first 2 months of life (Pope et al., 2005). While these data were limited in numbers, they suggest that human infants may only be more acutely sensitive to OP pesticides which are effectively inactivated by carboxylesterases during the early perinatal period. In addition to metabolic differences that may contribute to age-related sensitivity to OP pesticides, some toxicodynamic differences among age groups could also be important. Organophosphorus and carbamate pesticides are toxic by virtue of their ability to inhibit AChE (Fukuto, 1990). Species differences in sensitivity of AChE to inhibition by some OP anticholinesterases have been reported (Kemp and Wallace, 1990). Thus, there could be a degree of selective toxicity among age groups based on the molecular interaction between the toxicant and its “receptor,” AChE. Several studies have reported, however, that AChE sensitivity to the inhibitors is not a contributing factor to age-related differences in sensitivity (Atterberry et al., 1997; Benke and Murphy, 1975; Mortensen et al., 1998). It should be noted, however, that in the study by Pope and coworkers (2005) evaluating human liver carboxylesterase, chlorpyrifos oxon was significantly more potent at inhibiting activity in liver samples from a 2-month old infant, compared to activity in all other samples/ages. Thus, age-related differences in sensitivity of biotransformation enzymes to OP inhibitors could potentially contribute to differential sensitivity. Upon extensive inhibition of AChE in the nervous system, the neurotransmitter acetylcholine accumulates in synapses causing excessive stimulation of cholinergic receptors on postsynaptic cells leading to cholinergic toxicity. It is known that feedback inhibition of acetylcholine release can occur through activation of muscarinic acetylcholine receptors located on presynaptic terminals (Allgaier et al., 1993; Vickroy and Cadman, 1989; Weiler, 1989). Activation of these presynaptic muscarinic receptors diminishes further acetylcholine release and thereby may reduce the excessive stimulation of postsynaptic cholinergic receptors following extensive AChE inhibition. Pedata and coworkers (1983) reported that muscarinic autoreceptor function was absent in 7-day-old rat brain but viable in brain from 21-day-old animals. Thus, with extensive AChE inhibition, very young rats do not have an adaptive mechanism which limits further neurotransmitter release in times of excess (e.g., when AChE is inhibited). Pedata and coworkers (1983) and Meyer and Crews (1984) reported that evoked acetylcholine release was lower in tissues from both neonatal and aged brain compared to animals 1–6 months of age. Differences in the amount of acetylcholine released upon stimulation between the age

Hayes’ Handbook of Pesticide Toxicology

groups may therefore contribute to differences in response to AChE inhibitors. The function of muscarinic autoreceptors appears markedly reduced in some rat brain regions with aging (Araujo et al., 1990). Interestingly, Karanth and coworkers (2007) reported lesser striatal acetylcholine accumulation in aged rats compared to adults treated with parathion. A deficit or lack of feedback inhibition of acetylcholine release in some age groups, however, could limit their adaptation to synaptic AChE inhibition/acetylcholine accumulation and thereby contribute to higher sensitivity (Pope, 1999). Differences in acute sensitivity to OP anticholinesterases between neonatal and adult rats may therefore have both a toxicokinetic and toxicodynamic basis. It should be stressed, however, that the studies cited above generally used lethality as the endpoint for estimating age-related sensitivity. By definition, dosages at or near those causing death would have to be considered “high” level exposures. Less information is available regarding age-related differences in sensitivity to lower levels of exposure. While of prominent importance with acute, high level exposures where detoxification systems may be saturated, differential age-related detoxification capacities may have lesser importance when repeated, lower level exposures occur. With lower nonlethal dosages, less AChE activity would be inhibited with lesser signs of cholinergic toxicity. At even lower dosages, some degree of AChE inhibition could occur in the absence of any overt toxicity (Nostrandt et al., 1997). Under these conditions, feedback inhibition of acetylcholine release (or lack of that adaptive mechanism in neonatal animals) would have little consequence. Thus, with acute dosages of pesticide high enough to cause some level of AChE inhibition but with no alteration of cholinergic neurotransmission, two factors which could influence age-related differences in sensitivity (lower detoxification capabilities, lesser adaptive regulation of neurotransmitter release) may have no relevance. With repeated lower level exposures, however, another toxicodynamic factor (i.e., recovery of AChE activity following inhibition) may play a more prominent role. As mentioned before, AChE activity following OP exposure may recover much faster in neonatal tissues (Pope et al., 1991) and much slower in aged animals (Michalek et al., 1990) than in adult tissues. While neonatal rats were more sensitive to single, high dosages of chlorpyrifos, adults exhibited more extensive changes in cholinergic neurochemical markers (i.e., AChE inhibition, muscarinic receptor binding) following repeated, intermittent dosing (40 mg/kg, every 4 days for a total of four exposures) (Chakraborti et al., 1993; Pope and Liu, 1997). Apparently, while young rats are more sensitive to the acute effects of chlorpyrifos, they can recover much faster than adults to the biochemical insult. When exposures are separated in time sufficiently, neonatal animals can regain AChE activity faster and avoid cumulative inhibition with repeated exposures. By contrast,

Chapter | 32  The Influence of Age on Pesticide Toxicity

in particular with OPs such as chlorpyrifos which produce long-term inhibition of AChE, activity recovers more slowly in adult tissues allowing accumulative inhibition with subsequent exposures. Thus, under some conditions one can argue that with acute chlorpyrifos dosing, young animals are more sensitive than adults but with repeated dosing, age-related sensitivity is reversed. Clearly, the nature of the exposures (acute vs. repeated, high level vs. low level) can influence age-related differences in sensitivity to these toxicants. Relatively few studies have evaluated the effects of aging on sensitivity to organophosphates. Acetylcholine����������������������������������������������������������� sterase activity in some brain regions (e.g., hippocampus, cortex) but not others (e.g., pons-medulla) of rats declines with aging (Bisso et al., 1991; Meneguz et al., 1992). As mentioned before, recovery of AChE activity as well as muscarinic receptor binding following repeated organophosphate exposures was impaired in aging brain, in particular in cerebral cortex (Michalek et al., 1990). Karanth and coworkers (2007) reported that aged rats had significantly lower levels of total muscarinic receptor binding in striatum compared to adult rats. Age-related differences in baseline activity of cholinergic neurochemical processes or their adaptive responses to pesticide exposure could therefore influence sensitivity to some anticholinesterases. Veronesi and coworkers (1990) evaluated the effects of chronic fenthion exposure (25 mg/kg, three times a week for 10 months) in either young (2-month-old) or aged (12month-old) rats. Using this dosing treatment schedule, chronic (10 months) fenthion exposures initiated in young rats produced gliosis and necrosis in the dentate gyrus and CA4, CA3 and sometimes CA2 regions of the hippocampus. Aged rats treated with the same regimen of fenthion exhibited similar degrees of hippocampal degeneration earlier during the progression of exposure, i.e., by 2 months, and much more extensive pathology than noted in the younger animals when evaluated following 10 months of exposure. These studies suggest that persistent acetycholinesterase inhibition by fenthion can produce neuropathological changes in the rat hippocampus and that aged rats are more sensitive than younger rats to such effects. Karanth and Pope (2000) compared acute sensitivity to chlorpyrifos and parathion in neonatal (7-day-old), juvenile (21-day-old), adult (90-day-old) and aged (24-month-old) Sprague–Dawley rats. Neonatal and juvenile rats were more sensitive than adults to both toxicants. Adult and aged rats were similar in sensitivity to chlorpyrifos but aged animals were markedly more sensitive than adults to parathion. Moreover, plasma carboxylesterase activity among groups was highly correlated with acute sensitivity to parathion, further suggesting a toxicokinetic basis for the agerelated differences in sensitivity to this pesticide. Adults and aged rats treated with a range of dosages of parathion showed an approximately threefold difference in sensitivity based on striatal cholinesterase inhibition (Karanth et al.,

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2007). Interestingly, with dosages of parathion eliciting relatively similar degrees of cholinesterase inhibition (i.e., 27 mg/kg in adults and 9 mg/kg in aged rats), striatal extracellular acetylcholine levels were significantly lower in aged compared to adults rats (2.2- to 2.9-fold). While aged rats are more sensitive than adults to the acute toxicity of parathion, lesser CNS acetylcholine accumulation was noted in aged animals. The authors concluded that the lower density of muscarinic receptors noted in tissues from aged rats may influence the amount of acetylcholine accumulation required to elicit cholinergic signs. The above discussion pertains to differences in sensitivity among different age groups to the cholinergic toxicity of OP pesticides. A number of reports suggest that OP pesticides may affect macromolecular synthesis and cell viability in the brain following early postnatal exposures, independent of AChE inhibition (Slotkin, 1999). Whitney and coworkers (1995) reported that DNA and protein synthesis could be affected by chlorpyrifos in a time-dependent and brain regional-dependent manner. When postnatal rats (11–14 days of age) were given chlorpyrifos (1 mg/kg/day), a delayed reduction in DNA concentration and content in forebrain was noted at 15–20 days of age (Campbell et al., 1997). Reductions in cellular RNA concentration and content were also reported in brainstem and forebrain following repeated postnatal chlorpyrifos exposures in rats (Johnson et al., 1998). Song and coworkers (1997) reported that repeated postnatal exposures to chlorpyrifos in rats affected multiple components of the adenylyl cyclase cascade system (e.g., activity of adenylyl cyclase, G-protein function, expression of neurotransmitter receptors coupled to adenylyl cyclase). Moreover, changes in these processes were noted in cerebellum, a brain region with only sparse cholinergic innervation. More recent studies on neurodevelopmental toxicity suggested that targeting of the serotonergic system can occur at low level exposures to a number of OP insecticides (Aldridge et al., 2003, 2004; Moreno et al., 2008; Slotkin et al., 2006; Slotkin and Seidler, 2008). A number of epidemiological studies suggest that prenatal exposure to chlorpyrifos and possibly other organophosphorus insecticides, at levels far below those expected to elicit significant acetylcholinesterase inhibition, may lead to neurodevelopmental disturbances (Berkowitz et al., 2004; Perera et al., 2005; Whyatt et al., 2004). The specific macromolecular targets sensitive to OP toxicants that might contribute to these types of responses are unclear, however. See Chapter 33 for more information about longterm functional consequences of developmental exposure to these types of pesticides. Anticholinesterases may affect neuronal adhesion and neurite extension, possibly by direct binding to noncatalytic sites of the enzyme (Bigbee et al., 1999; Dupree and Bigbee, 1994; Small et al., 1995; Song et al., 1998). Blasina and coworkers (2000) reported that the acetylcholinesterase peripheral binding site ligand fasciculin modified chicken

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retinal development in vitro at concentrations that did not affect acetylcholinesterase catalytic activity. Coating of tissue culture plates with acetylcholinesterase stimulated neurite outgrowth in neuroblastoma cells, and this response was inhibited by peripheral site antagonists but not by the active site inhibitor, eserine (Johnson and Moore, 2000). Peripheral site antagonists were hypothesized to promote neurite extension/remodeling in hippocampal neurons via induction of RACK-1 (receptor for activated C kinase) signaling (Farchi et al., 2007). Using dorsal root ganglion neurons from AChE/ and / mice, Yang and colleagues (2008) showed that acetylcholinesterase was essential for chlorpyrifos-induced decreases in axonal length occurring at concentrations insufficient to inhibit catalytic activity. Thus, OP pesticides may be capable of altering macromolecule synthesis, intracellular signaling and neuronal adhesion/outgrowth in the developing brain, apparently independent of catalytic inhibition of AChE. A number of other serine active hydrolases can be targets for OP anticholinesterases (Casida and Quistad, 2004, 2005). Enzymes that degrade endocannabinoids (e.g., monoacylglycerol lipase, fatty acid amide hydrolase) have been shown to be quite sensitive to inhibition by a variety of OP inhibitors (Quistad et al., 2002, 2006). The high sensitivity of these enzymes to many OP inhibitors relative to acetylcholinesterase, and the role of endocannabinoid signaling in modulation of neurotransmission (Kano et al., 2009), suggest these noncholinesterase targets may have a toxicologically relevant role in expression of cholinergic toxicity (Nallapaneni et al., 2006, 2008; Quistad et al., 2006). While the targeting of serine active sites in target proteins has been known for decades, more recent evidence suggests organophosphorylation of tyrosine residues could be important in acute and persistent consequences of anticholinesterase exposures (Grigoryan et al., 2008; Li et al., 2009; Williams et al., 2007). Little is known, however, regarding whether these additional macromolecular targets may be involved in age-related differences in sensitivity to OP anticholinesterases. Some organophosphorus toxicants can induce a delayed neuropathological disorder referred to as organophosphorusinduced delayed neurotoxicity (OPIDN) (Abou-Donia, 1981). This form of neurotoxicity is not associated with AChE inhibition but has been correlated with the inhibition of another enzyme in the nervous system called neurotoxic esterase (NTE) (Johnson, 1976, 1980). Individuals affected by this delayed neurotoxicity exhibit gait disturbances (incoordination and difficulties in walking) and sensory deficits (numbness and tingling, particularly in the fingers and toes), which may or may not follow signs of toxicity characteristic of AChE inhibition. Degeneration of certain nerve tracts in both the central and peripheral nervous systems has been demonstrated in OPIDN. It has more recently been observed that some compounds [e.g., the common protease and NTE inhibitor phenylmethylsulfonyl

Hayes’ Handbook of Pesticide Toxicology

fluoride (PMSF)], while not being capable of inducing delayed neurotoxicity can potentiate or promote delayed neurotoxicity caused by an OP (Lotti et al., 1991; Pope and Padilla, 1990; Pope et al., 1993). The sequence of administration of the two compounds is of paramount importance, i.e., for delayed neurotoxicity to be exacerbated, OP exposure must precede exposure to the potentiating agent. Young animals are resistant to delayed neurotoxicity (Johnson and Barnes, 1970; Moretto et al., 1991). Before the age of about 6–7 weeks, chickens (the animal model of choice for studies of delayed neurotoxicity) are completely resistant to functional and morphological signs of OPIDN. From about 7–10 weeks of age, sensitivity develops and at about 12–14 weeks of age, they become completely sensitive (Moretto et al., 1991; Pope et al., 1992). Studies have also examined the potentiation of OPIDN in young animals (Peraica et al., 1993; Pope et al., 1992). As stated above, 5-week-old chickens are normally resistant to the clinical and morphological changes associated with delayed neurotoxicity. If OP exposure is followed by treatment with PMSF, however, overt delayed neurotoxicity can be demonstrated. On the other hand, clinical and morphological changes typical of OPIDN are generally not elicited in very young chickens (e.g., 2 weeks of age) regardless of the dose of the OP or whether a potentiating agent is given after the OP (Harp et al., 1997). Just as the mechanism(s) underlying OPIDN itself has not been elucidated, the basis for age-related differences in sensitivity to delayed neurotoxicity remains unknown. In contrast to age-related sensitivity to acute toxicity from most OPs, however, young animals are less sensitive than adults to the delayed neurotoxicity of OPs.

32.3.2  Organochlorine Insecticides At one time, organochlorines (OCs) constituted the highest-use pesticide class in the world. With increased awareness of ecological damage, global contamination, and insect resistance, the use of OCs has decreased. The most well-known OC, DDT has been extensively studied. In acute toxicity studies, DDT is actually less toxic to neonatal rats than to adults (Lu et al., 1965). In this same study, dieldrin, another OC, was also reported to be less toxic in neonatal rats. Several studies have suggested that early neonatal exposure to DDT (0.5 mg/kg, po) can have long-lasting consequences (Eriksson et al., 1984, 1993). Total cholinergic muscarinic receptor ([3H]QNB) density was increased in cortex 1 week after DDT exposure in 10-day-old rats but no effect was noted in hippocampus. Moreover, muscarinic receptor binding was still altered at 4 months of age following this single treatment with DDT, but at this time there was a reduction in binding density. Functional alterations (deficits in locomotor habituation) were also noted in rats 4 months after acute DDT exposure (Eriksson, 1997).

Chapter | 32  The Influence of Age on Pesticide Toxicity

Of particular interest in these studies was the observation that neonatal (10-day-old) rats treated with DDT (0.5 mg/kg) showed an increase in cortical muscarinic receptor binding 1 week after exposure whereas adult rats treated similarly showed a decrease in receptor binding. Moreover, neither 3-day-old rats nor 19-day-old rats showed the same response (i.e., upregulation of muscarinic receptors) when treated similarly with DDT (Eriksson, 1997). Subsequent studies have shown that 10-day-old mice treated with DDT and then challenged at 5 months of age with bioallethrin showed increased expression of the m4 subtype of muscarinic receptors in selected brain regions (cortex and striatum) (Talts et al., 1998a). Thus, there appears to be a critical developmental window in which alteration of the cholinergic system can occur following early DDT exposure, and changes in muscarinic receptor density induced by DDT appear specific for the m4 subtype. While most OCs have been banned from use in the United States, their use continues in other countries. Endosulfan is still registered for about 80 crop uses in the United States. Lindane (-hexachlorocyclohexane) use was voluntarily cancelled in the United States in 2006, with all remaining pesticidal products to be used by 2009. Lindane is still commonly prescribed, however, for treatment of scabies and pediculosis. Rivera and coworkers (1990) reported that repeated, relatively low-level exposures to lindane (10 mg/kg/day for 7 days) during postnatal week 1 or 2 induced transient changes in reflex behaviors (e.g., surface righting, cliff avoidance) and locomotor hyperactivity, in the absence of overt signs of toxicity. Serrano and coworkers (1990) reported that early postnatal lindane exposure reduced the level of myelin basic protein and 2,3-cyclic nucleotide 3-phosphodiesterase activity, an enzyme in high concentrations in myelin and myelin-forming cells, in a dose-dependent manner. Lindane exposure (either acute [20 mg/kg] or repeated [10 mg/kg/day for 7 days]) in rats 15 days of age caused complex behavioral changes (improvement in passive avoidance behavior, alterations in locomotor activity) and apparent enhanced turnover of brain monoaminergic neurotransmitters (Rivera et al., 1998). While these studies only evaluated toxicity in postnatally maturing animals, the endpoints evaluated and the changes noted suggested that higher sensitivity may exist in younger individuals. Samanta and Chainy (1997) reported that acute lindane exposure (50 mg/kg, i.p.) caused only minimal lipid peroxidation in liver of 30-day-old chickens but more extensive oxidative changes in 7-day-old animals. Furthermore, superoxide dismutase was inhibited and glutathione levels were elevated by lindane in 7-day-old but not 30-day-old chickens. Thus, lindane can cause diverse age-related changes that generally target younger animals. Kiran and Varma (1988) studied the toxicity of endosulfan in different age groups of rats (12.5 mg/kg/day for 4 days beginning at 15, 30, 70 and 365 days of age).

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Hyperglycemia and glycogen depletion were most extensive in 356-day-old animals and least affected in the youngest age group. Liver aldolase activity was also reduced more in older rats than in younger animals. By contrast, red blood cell Na/K ATPase activity was inhibited more in the youngest age group. Jia and Misra (2007) reported that postnatal exposure (0.155 mg/kg/day from PND 5–19) in mice to endosulfan with subsequent challenge at 8 months of age significantly decreased striatal dopamine and DOPAC levels. These results suggest complex age-related differences in response to endosulfan. Chlordecone is an organochlorine that causes hyperexcitability, tremors, incoordination, and other signs of neurotoxicity (Tilson and Mactutus, 1982). Several studies evaluated the effects of early postnatal exposure to chlordecone. Tilson and coworkers (1982) reported that rats exposed acutely on postnatal day 4 to chlordecone had markedly altered responses during reversal of visually cued nose poke behavior when tested at about 4 months of age. Neonatal chlordecone exposure was also reported to alter passive avoidance performance (Mactutus et al., 1982). Jinna and coworkers (1989) reported that chlordecone inhibited rat brain ATPases (Na/K ATPase, Ca ATPase) in an age-related manner, i.e., neonatal enzyme activity was more sensitive to inhibition by chlordecone in vitro. Chlordecone has been shown to potentiate the hepatotoxicity of halogenated solvents, e.g., carbon tetrachloride (Soni and Mehendale, 1998). Rats of ages 20 and 45 days were resistant to chlordecone-enhanced hepatotoxicity relative to 60-day-old animals, however (Dalu and Mehendale, 1996). Dosages of chlordecone (10 ppm in the diet for 15 days) and carbon tetrachloride (0.1 ml/kg, i.p.) that caused 100% lethality in the adult rats caused 0% and 25% lethality in 20- and 45-day-old animals. It was concluded from these studies that the relative ability of the liver to recover from injury was the prominent factor underlying age-related differences in toxic outcome, with immature animals being more competent than adults at restoring tissue integrity and function. Thus, while these studies do not indicate age-related differences in sensitivity to chlordecone alone, they suggest that the modulation of solvent hepatotoxicity by chlordecone can occur in an agerelated manner. Many of the OCs, e.g., DDT, chlordecone, methoxychlor, chlordane, and endosulfan, have also been noted to interact directly with hormonal receptors (Tilson, 1998). The DDT analog, methoxychlor, only recently removed from use in the United States, has been shown to both alter sex-related hormones and reproductive function in rats treated postnatally (Chapin et al., 1997). The endocrine-disrupting capa­ city of these agents could be cause for concern with early exposures (Davis et al., 1993; Cassidy et al., 1994; Chapin et al., 1997). The reader is referred to Chapter 18 for more information on endocrine disruption by pesticides.

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32.3.3  Carbamates Knisely and Hamm (1989) investigated the comparative actions of physostigmine on nociception in different age groups of rats (3, 17 and 25 months of age). Tail-flick latencies were dose-dependently altered in all age groups by physostigmine, but more extensive increases in latency were noted in the 17- and 25-month-old animals with higher dosages, suggesting higher sensitivity in the aged animals to this carbamate anticholinesterase. Such changes could be an indication of upregulation of cholinergic receptors due to loss of cholinergic innervation with aging. Takahashi and colleagues (1991) compared the motor, sensory and thermoregulatory responses of young adults (3 months of age) and older adults (12 months of age) to carbaryl (10 or 50 mg/kg, p.o.). Carbaryl affected nociception primarily in the older animals. Hypothermia was also affected in an age-related manner. Locomotor changes following carbamate exposure, however, were similar between the two age groups. Again, these data illustrate the potential for age-related differences in response to a pesticide when one endpoint is used and, conversely, lack of age-related differences in sensitivity when based on another endpoint. Moser (1999) reported that aldicarb was about twice as toxic in preweanling rats compared to adults using the acute maximum tolerated dose as the endpoint of sensitivity. Interestingly, preweanling rats exhibited fewer signs of functional toxicity than older animals, in the presence of similar levels of brain and blood cholinesterase inhibition. Furthermore, the young rats were resistant to locomotor alterations noted in older animals following aldicarb administration. Lifshitz and coworkers (1997) retrospectively compared the clinical course of poisoning in children (1–8 years of age) and adults (17–41 years of age) following carbamate pesticide exposures. In all cases, blood serum cholinesterase inhibition was approximately the same (10–30% below the lower limit of normal). Interestingly, signs of coma/ stupor and hypotonia were noted in 100% of the children but in none of the adults. While miosis was noted in 92% of the adults, this sign was only recorded in 55% of the children. Moreover, muscle fasciculations were observed in 83% of the adults and in only 6% of the children. While the relative level of AChE inhibition in the target tissues was unknown, these results suggest that children may respond differently than adults following acute anticholinesterase exposures producing relatively similar degrees of blood cholinesterase inhibition.

32.3.4  Pyrethroid Insecticides Eriksson and Nordberg (1990) studied the effects of early postnatal exposures to one of two different pyrethroid insecticides, bioallethrin (a type I pyrethroid) and deltamethrin (a type II pyrethroid), on cholinergic receptors in

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mouse brain. With lower levels of exposure, bioallethrin (0.72 mg/kg/day from postnatal days 10 to 16) reduced high affinity muscarinic receptor binding in brain whereas deltamethrin (0.71 mg/kg/day) increased high affinity binding, both in the absence of overt signs of toxicity. Deltamethrin also increased cortical [3H]nicotine binding. Higher levels of repeated exposure (72 and 1.2 mg/kg/day for bioallethrin and deltamethrin, respectively) caused overt toxicity (tremor, choreoathetosis) but only deltamethrin affected cholinergic receptor binding under these conditions. Early exposure to bioallethrin in mice (0.7 mg/ kg/day from postnatal days 10 to 16) was also reported to increase sensitivity to bioallethrin when administered at 7 months of age, suggesting long-term changes in sensitivity following exposure during postnatal maturation (Talts et al., 1998b). These studies, similar to those with early postnatal exposures to DDT (Eriksson et al., 1984), indicate that development of some components of the cholinergic system may be sensitive to alteration by early postnatal exposure to “noncholinergic” pesticides (i.e., pesticides not having a primary action on some aspect of cholinergic neurotransmission). Cantalamessa (1993) compared the acute toxicity and metabolism of cypermethrin and permethrin in neonatal and adult rats. With both pesticides, an age-related decrease in acute toxicity was noted. Cypermethrin and permethrin were 16.8 and 4.4 times more toxic (based on 24-hour oral LD50 values) in 8-day-old animals compared to adults. Carboxylesterase inhibition (by tri-ortho-cresyl phosphate) in neonatal animals failed to alter acute toxicity but lethality was increased in adults by this pretreatment, suggesting that neonatal animals may be more sensitive to acute toxicity of these pyrethroids at least partially because of incomplete development of this detoxification system. Sheets and coworkers (1994) evaluated the sensitivity of preweanling, weanling and adult rats to a wide dose range of deltamethrin. Younger rats (11 and 21 days of age) were markedly more sensitive than adults (72 days of age) to the acute lethality of deltamethrin (LD50: 11 days  5.1 mg/kg; 21 days  11 mg/kg; 72 days  81 mg/kg, p.o.). By contrast, using acoustic startle response to evaluate functional toxicity of lower level exposures, the ED50 was the same between 11-day-old and 72-day-old animals. Based on these studies, age-related differences in sensitivity to deltamethrin could be considered substantial (if based on acute lethality) or nonexistent (if based on the acoustic startle response). Clearly, the selection of the endpoint used to define sensitivity as well as the exposure conditions can qualitatively influence determination of age-related susceptibility to these pesticides.

32.3.5  Miscellaneous Pesticides Gaines and Linder (1986) examined the comparative acute sensitivity of weanling (4–6 weeks of age) and adult

Chapter | 32  The Influence of Age on Pesticide Toxicity

rats to 34 pesticides from different chemical classes. The immature rats were more sensitive to only four of those pesticides. Moreover, differences in acute sensitivity to pesticides were generally only two- to three-fold in magnitude. One problem with this study, however, was the age of the younger animals used, i.e., 4- to 6-week-old rats. Similar studies using less mature animals (e.g., 1- to 3-week-old rats) may have yielded different conclusions. Watkinson studied the cardiotoxicity of the formamidine pesticide chlordimeform. Weanling (22–30 days of age; Watkinson, 1985) and aged (24 months of age; Watkinson, 1986) rats were treated sequentially with 5, 10, 30, 60 and 120 mg/kg chlordimeform (i.v.) or vehicle and mean arterial blood pressure and heart rate monitored. While chlordimeform reduced heart rate and blood pressure in both age groups, the magnitude of the changes was greater in the aged animals. Arrhythmias were also less pronounced in younger animals and required higher thresholds of chlordimeform. In addition, while a single injection of chlordimeform (60 mg/kg, i.v.) was lethal to all aged rats tested, only 23% of the weanling rats died following this level of exposure. Lower sensitivity of young rats to the lethality of chlordimeform had been previously reported (Robinson and Smith, 1977). Thus, it appears that younger animals are less sensitive than aged rats to the toxicity of the formamidine insecticide, chlordimeform. Ivermectin is a broad spectrum antiparasitic agent (Campbell and Benz, 1984). Relatively low level exposure to ivermectin (4 mg/kg/day) during gestation (GD 6–20) and lactation (postnatal days 2–20) caused 100% lethality in pups with no apparent toxicity in dams (Poul, 1988). When exposure was limited to gestation, only 22% lethality was noted in the offspring. Lower exposure levels (1 mg/kg/day) had no effect on survival but delayed some developmental endpoints including cliff avoidance and locomotion. Lankas and coworkers (1989) reported that newborn rodents were particularly sensitive to the neurotoxicity of ivermectin. Following application of ivermectin to control an ectoparasite infestation, Skopets and coworkers (1996) noted evidence of higher sensitivity of young mice to ivermectin. While all adults tolerated the ivermectin exposures, preweanling mice developed seizures or tremors and lethality was observed in some cases. Together, these data suggest that neonatal rodents are more sensitive than adults to the acute toxicity of ivermectin. As ivermectin is typically prevented from access to the central nervous system in adults (Lovell, 1990), incomplete blood–brain barrier formation appears to contribute to these age-related differences in sensitivity (Lankas et al., 1989). Age-related differences in sensitivity were noted following acute dibromochloropropane exposure (250 mg/kg, s.c.) in 4- and 9-week-old rats (Saegusa, 1987). It was noted that the older animals exhibited a higher incidence of lethality, more extensive body weight reductions, and more extensive tissue damage in kidney, intestine, and testes.

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Dithiobiuret (DTB, thioimidodicarbonic diamide) was originally proposed as a rodenticide and is a prototypical motor neuron toxicant that produces a flaccid weakness following repeated exposures (Atchison et al., 1982). Using failure of the rotorod test as an indication of neuromuscular toxicity, Atchison and coworkers (1982) studied the sensitivity of weanling (25-day-old), juvenile (50-day-old) and adult (80-day-old) rats to DTB (1 mg/kg/day). In females, the mean time to onset of rotorod failure was about 6 days in weanling rats, 4 days in juveniles, and only about 3 days in adults. Neither differences in total accumulation of DTB nor distribution appeared to contribute to the differences in DTB toxicity among the age groups. These data provide another example of higher sensitivity to neurotoxicants in adults compared to younger animals. Using a series of immunotoxicity assays, Smialowicz and coworkers (1989) reported that preweanling rats (3–24 days of age) were somewhat more sensitive than adults to tributyltin-induced immune alterations. In addition, natural killer cell activity was only affected in the neonatal animals. Furthermore, some immune responses were altered in 10-week-old animals treated prior to weaning, suggesting long-term changes in immune function could occur with early exposure to tributyltin. Children may be more sensitive to the insect repellant, DEET (diethyl m-toluamide) (Couch and Johnson, 1992). DEET is used safely by an estimated 200 million people each year around the world (Brown and Hebert, 1997) but severe neurological manifestations have occasionally been associated with its use (Osimitz and Murphy, 1997). Four boys (age 3–7 years) had seizures within 48 h of applying DEET to the skin. Six young girls (ages 1.5–8 years) exhibited seizures, ataxia and/or coma after dermally applying DEET and three of those children later died. These types of neurological signs have been reported in adults following oral consumption of large amounts of DEET (Tenenbeim, 1987). Thus, while rare in occurrence, children may exhibit serious signs and symptoms of neurotoxicity and can die following dermal application of this widely used repellant. Because of the scarcity of data on absorption, metabolism or elimination of DEET in children, it is unclear why children may be more sensitive to this compound (Garrettson, 1997). Use of fipronil, a commonly used insecticide, is increasing because of the loss of other products, e.g., OPs. Some studies suggest that the developing nervous system may be particularly sensitive to fipronil. Stehr and coworkers (2006) reported neurodevelopmental effects in zebrafish, possibly working through glycine receptor interactions. A more recent in vitro study suggests that fipronil may have a host of developmental effects including DNA and protein synthesis inhibition, induction of oxidative stress, and reduced cell density in a PC12 cell model (Lassiter et al., 2009). While the target macromolecule for fipronilinduced changes was not evaluated in these studies, PC12

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cells lack the GABAA receptor typically considered the target for phenylpyrazole insecticides. Thus, developing organisms may be sensitive to such effects.

Conclusion Changes in sensitivity to pesticides can occur throughout the lifespan from early postpartum to senescence. Recently, there has been considerable concern that children may be at higher risk than adults to pesticides. Enactment of the Food Quality Protection Act in 1996 was in response to this concern, calling for consideration of additional safety in the risk assessment of pesticides to protect infants and children. It is clear from review of both experimental and clinical data, however, that there is no hard-and-fast rule regarding age-related differences in sensitivity to pesticides. While neonates may be more sensitive to the acute toxicity of some pesticides, adults may be more sensitive to others. Even within a class of toxicants, e.g., within the organophosphorus anticholinesterases, examples of higher sensitivity in both age groups can be demonstrated. In fact, even when a single pesticide is considered, age-related differences in sensitivity may change qualitatively depending on the conditions of exposure (e.g., acute vs. repeated dosing, high vs. low level exposures) or the endpoint measured. While maturational differences in biotransformation capacity may be limiting in some cases, e.g., with acute, high level exposures where detoxification enzymes could become saturated, such metabolic differences may be of lesser importance with repeated, lower levels of exposure to the same pesticides. Similarly, differences in the ability to recover following pesticide exposure may be much more important when repeated exposures occur than following acute exposures. Storage and clearance of pesticides may also be more important with repeated, long-term exposures. Age-related sensitivity to pesticides should therefore be evaluated on a case-by-case basis, recognizing both the factors which influence age-related differences in response and the critical importance of appropriate endpoint selection for establishing differential sensitivity. Relative sensitivity can be expressed in one of two ways, i.e., a subpopulation exhibits differences in sensitivity to a particular form of toxicity or a subpopulation exhibits qualitatively different forms of toxicity to the same pesticide. Young animals may be more sensitive to the acute lethality of some pesticides but this does not necessarily mean that young animals will be more sensitive to the same pesticides when sensitivity is based on nonlethal endpoints of toxicity. Risk assessments are typically performed using a “critical” endpoint, generally the most sensitive endpoint to the toxicant in question derived from a series of toxicity studies. Thus, even if a pesticide causes a qualitatively different form of toxicity in different age groups, the risk assessment and estimation of tolerable exposure levels will

not change unless this response occurs at dosages lower than those defining the critical endpoint. There will always be uncertainty in risk assessment. One factor which contributes to that uncertainty is age and its influence on the response to a particular toxicant. If the critical endpoint for a particular pesticide is well established based on “reliable” data derived from studies across all age groups, the contribution of age-related differences in sensitivity to such uncertainty can be minimized. Knowledge of mechanisms which contribute to such agerelated differences in response to pesticides will ultimately aid in the safer use of these chemicals.

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The Influence of Age on Pesticide Toxicity

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administration in the postweanling rat. J. Toxicol. Environ. Health 19, 207–217. Weiler, M. H. (1989). Muscarinic modulation of endogenous acetylcholine release in rat neostriatal slices. J. Pharmacol. Exp. Ther. 250, 617–623. Wennberg, R. P. (1993). Animal models of bilirubin encephalopathy. Adv. Vet. Sci. Comp. Med. 37, 87–113. Whitney, K. D., Seidler, F. J., and Slotkin, T. A. (1995). Developmental neurotoxicity of chlorpyrifos: cellular mechanisms. Toxicol. Appl. Pharmacol. 134, 53–62. Whyatt, R. M., Rauh, V., Barr, D. B., Camann, D. E., Andrews, H. F., Garfinkel, R., Hoepner, L. A., Diaz, D., Dietrich, J., Reyes, A., Tang, D., Kinney, P. L., and Perera, F. P. (2004). Prenatal insecticide exposures and birth weight and length among an urban minority cohort. Environ. Health Perspect. 112, 1125–1132.

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Wilkinson, C. F., and Ginevan, M. E. (1989). A critical review of the Natural Resources Defense Council’s Report. In “Intolerable Risks: Pesticides in Our Children’s Food”. RiskFocus, Versar, Inc., Springfield, VA. Williams, N. H., Harrison, J. M., Read, R. W., and Black, R. M. (2007). Phosphylated tyrosine in albumin as a biomarker of exposure to organophosphorus nerve agents. Arch. Toxicol. 81, 627–639. Wynne, H., Mutch, E., James, O. F., Rawlins, M. D., and Woodhouse, K. W. (1987). The effect of age on mono-oxygenase enzyme kinetics in rat liver microsomes. Age Ageing 16, 153–158. Yang, D., Howard, A., Bruun, D., Ajua-Alemanj, M., Pickart, C., and Lein, P. J. (2008). Chlorpyrifos and chlorpyrifos-oxon inhibit axonal growth by interfering with the morphogenic activity of acetylcholinesterase. Toxicol. Appl. Pharmacol. 228, 32–41.

Chapter 33

Lasting Behavioral Consequences of Organophosphate Pesticide Exposure During Development Olga A. Timofeeva and Edward D. Levin Duke University Medical Center, Durham, North Carolina

33.1  Introduction

33.2  Human effects

Impaired neurobehavioral development of children has been significantly linked in epidemiological studies with exposure to pesticides. Lowering pesticide exposure decreased the degree of impairment. Environmental epidemiological studies are essential for identifying toxic risks, but like any single scientific approach they have limitations. Determining the cause-and-effect relationship beyond significant association is a challenge. Defining exposure to particular chemicals within a class is often difficult. Prospectively tracing early developmental exposure to persisting effects in adulthood and aging can take great lengths of time. Finally, epidemiology is the study of damage done. The aim of toxicology is to be a predictive science, to prevent toxic damage. Animal model studies can work in concert with epidemiological research to resolve many of these issues. With regard to organophosphate (OP) pesticide-induced developmental neurobehavioral toxicity, laboratory animal model studies have clearly demonstrated that the developing nervous system is quite vulnerable to detrimental effects of OP pesticides, even if exposure was short-term and at doses that did not cause much inhibition of acetylcholinesterase. Animal studies revealed that developmental OP exposure affects emotional and cognitive functions, social responses, and sex-related behavioral patterns. Genetic vulnerabilities have been identified. Neurobehavioral impairments have been shown in adults long after short-term exposure to OPs in pre- or postnatal periods. Animal studies support the idea that OP pesticides are a serious danger to children’s health and development.

It has been conclusively shown that OP pesticides can exert significant adverse neurotoxic effects in nontarget species, including humans. Because of the phosphorylation of acetylcholinesterase (AChE), they cause cholinergic toxicity. Given that insect and mammalian nervous systems both involve cholinergic mechanisms to perform vital functions, these compounds are responsible for the thousands of poisonings and deaths occurring annually as a result of pesticide exposures worldwide. A few OPs can also cause another type of toxicity known as organophosphateinduced delayed polyneuropathy (OPIDP), which can be classified as distal sensorimotor axonopathy. OPIDP is not related to AChE inhibition. Extensive studies carried out during the past 30 years have identified another esterase, called neuropathy target esterase (NTE), as a primary target for OPIDP (Lotti and Moretto, 2005). Data on subjects acutely poisoned with organophosphorus compounds show long-lasting impairment in neurobehavioral performance (reduction of verbal attention, memory, visual attention, flexibility of thinking) and, in some cases, impairments of emotional function. These impairments could be caused by either direct cholinergicmediated or noncholinergic neurotoxicity (Colosio et al., 2003). Identifying the critical mechanisms of OP-induced neurotoxicity would significantly improve our knowledge of the neurobehavioral effects of OPs. While the acute effects and targets of high-dose OP exposure have been, for the most part, clearly identified and characterized by numerous animal studies and cases

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of human poisoning, the effects of low-dose chronic or short-term OP exposure on human health and behavior are not well enough understood (Colosio et al., 2003; Daniell et al., 1992; Fiedler et al., 1997; Jamal et al., 2002; London et al., 1997; Ray and Richards, 2001). Low-level OP exposure can be defined as exposure that does not evoke cholinergic symptoms such as lacrimation, salivation, meiosis, or muscle fasciculation (Ray and Richards, 2001). Among the symptoms frequently reported in workers chronically exposed to low doses of pesticides are the following: anxiety, impaired vigilance, reduced memory and learning, fatigue, and reduced attention (Colosio et al., 2003). Very low exposures to environmental toxicants may lead to diseases that resemble many common illnesses that have other causes, or they may lead to decrements of functioning that are subtle and nonspecific. Determining a link between human exposure to the low dose of a specific chemical and long-term expression of a change in human health or behavior constitutes a tremendous challenge when designing an epidemiological study. Such studies have similar limitations, mainly in the definition of levels of exposure and in the selection of control groups (Colosio et al., 2003). In addition, functional impairments, particularly behavioral, can be difficult to completely determine. They require a variety of measures to detect the spectrum of abnormality and its extent. Thus, a battery of tests is continuing to evolve to measure with increasing sensitivity psychomotor, psychological, clinical, and psychiatric symptoms to better quantify functional impairment. In addition, neurophysiological tests are usually deployed in complex circumstances in which many factors, including economic status and education, combine to produce a particular effect such as lowered intelligence quotient score. Moreover, some consequences of early damage may not even emerge until advanced age. Only well-planned, sophisticated epidemiologic and animal studies can answer the questions that pertain to the toxicity of low-level exposures to environmental toxicants. Several elegant epidemiological studies suggested that chronic low-dose exposure to pesticides could be associated with increased risk for Parkinson’s disease (Barbeau et al., 1987; Langston, 1998; Ritz and Yu, 2000). Of particular interest is a recent investigation by Ascherio et al. (2006), which was conducted on a large (comprising more than 140,000 participants) cohort of men and women living in the United States. The study revealed that individuals exposed to pesticides had a 70% higher incidence of Parkinson’s disease than those not exposed. In another recent study (Berkowitz et al., 2004), the effects of low-dose pesticide exposure of mothers, recruited from East Harlem and other sections of New York City, on their infant’s birth weight, length, and head circumference was evaluated. The low levels of pesticide exposure (based either on questionnaire responses or the level of maternal urinary pesticide metabolites) were not found to

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affect fetal growth or gestational age. However, a small but significant impairing effect of chlorpyrifos was revealed when the level of maternal paraoxonase, an enzyme that detoxifies OP chlorpyrifos before it can inhibit acetylcholinesterase, was taken into account. The study found that low maternal paraoxonase activity coupled with chlorpyrifos levels just above the limits of detection, were associated with a small reduction in infant’s head circumference. Head circumference has been shown to correlate with brain weight (Lemons et al., 1981). Both brain size and head circumference are, in turn, predictive of cognitive ability (Ounsted et al., 1988). This study demonstrated that even tiny concentrations of chlorpyrifos could be detrimental for human brain development and that the activity of the human enzyme paraoxonase needs to be considered in evaluation of vulnerability to OP pesticide exposure. A positive correlation between exposure to OP pesticides and newborn head circumference was also found in a study conducted on 200 pregnant mothers living in agricultural areas in Argentina (Souza et al., 2005). The authors also reported that placental enzymes, such as AChE, gluthianone S-transferase, and catalase, may be used as biomarkers of prenatal exposure to OPs. Another recent study (Whyatt et al., 2004) reported an association between umbilical cord plasma chlorpyrifos levels and fetal birth weight decreases among minority women living in New York City during pregnancy. The study raised a debate as to whether impaired fetal development could be a critical noncholinergic effect rather than the inhibition of AChE (Zhao et al., 2005). Garry and colleagues, in a large (1,532 subjects) human epidemiologic study, uncovered weak but statistically significant associations between neurodevelopmental impairments and exposure to pesticides (Garry et al., 2002). In a more recent study (Young et al., 2005), which involved 381 infants and their mothers, living in the agricultural community of the Salinas Valley, California, an association was found between in utero OP exposure (which was determined by urinary levels of nonspecific OP dialkylphosphonate metabolites) and abnormal reflexes in neonates. The same cohort was investigated in a follow-up study (Eskenazi et al., 2007), which reported a negative association of prenatal and a positive association of postnatal OP pesticide exposure (level of dialkylphosphonate metabolites) with the mental development of infants evaluated at 24 months of age. This study highlights the possible importance of developmental stage on the neurotoxic outcome of OP pesticide exposure. In a smaller cohort of 254 children living in an inner-city minority population, Raugh et al. (2006) investigated the neurotoxic effects of prenatal chlorpyrifos (CPF) exposure on their cognitive and motor development at 12, 24, and 36 months of age. With the higher CPF exposure levels, they found a significant increase in the proportion of children with delayed psychomotor and mental development and children with attentional deficits and pervasive developmental disorder. The

Chapter | 33  Lasting Behavioral Consequences of Organophosphate Pesticide Exposure During Development

negative effects of CPF exposure were seen even when they included factors such as environmental tobacco smoke, maternal IQ, and maternal education status as covariates in the analysis, supporting the significant association of neurobehavioral impairment with CPF exposure. Specific physiological aspects of the developing organism, particularly in the CNS, may render it more susceptible to potential toxicants, including both toxicokinetic and toxicodymamic components (Faustman et al., 2000). Additional mechanistic information is needed. Animal model research can be particularly helpful (Brent and Weitzman, 2004; Weiss et al., 2004). Pesticides are ubiquitous in the environment. They are found in food, water, homes, schools, workplaces, and lawns (Lu et al., 2004; Morgan et al., 2005; Quandt et al., 2004). The unique childhood behaviors and activities such as hand-to-mouth activity place them at greater risk for heavier exposure to contaminants such as pesticides compared with adults (Weiss et al., 2004). Childhood neurobehavioral development as a target for environmental toxicants is a specific issue and a continuous concern in public health (Tilson, 2000). The development of noninvasive sampling methods, such as testing pesticides and their metabolites in urine, has made it possible to monitor pesticide exposure not only in mothers, but also in infants and children. The results of such monitoring contributed evidence that every child conceived today in the northern hemisphere is exposed to pesticides from conception throughout gestation and lactation regardless of where it is born (Colborn, 2006). Animal models are valuable to gain a better understanding of the risks for humans from OP-induced neurotoxicity, the mechanisms for these adverse effects, and ways to minimize such risk.

33.3  Rat models The mechanistic bases for human neurodevelopmental toxicity from OPs is currently being investigated in animal models. An increasing body of literature has demonstrated that developmental exposure to OPs, at doses causing little or no inhibition of AChE, results in neurochemical (Aldridge et al., 2003; Qiao et al., 2003; Slotkin and Seidler, 2007; Slotkin et al., 2006b, 2008c) and behavioral abnormalities (Icenogle et al., 2004; Levin et al., 2001, 2002; Roegge et al., 2008; Timofeeva et al., 2008a) in experimental animal models in which cause-and-effect relationships can be proved. Slotkin and colleagues discovered that important aspects of low-dose OP toxicity are not the result of AChE inhibition, but of other numerous mechanisms that alter the development and function of a number of regions of the brain (Aldridge et al., 2003, 2005b; Garcia et al., 2001; Roy et al., 2004; Slotkin et al., 2008a). Thus, recent animal studies clearly demonstrate that short-term, low-dose exposure to OP pesticides cause

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lasting neurobehavioral impairment. Further animal studies are needed to better model human conditions, such as exposure to mixed OP pesticides and industrial pollutants, and chronic exposure. The development of more sensitive neurobehavioral tests could be of particular importance in reducing uncertainty about the nature of adverse effects of developmental exposure to environmental neurotoxicants. Studies undertaken during the last few years consistently demonstrated that developing rat brain is highly vulnerable to low doses of OP pesticide exposure, even if exposure lasted only a few days and did not cause overt signs of systemic toxicity or significant AChE inhibition. The window of such vulnerability is wide and spans from at least gestational day (GD) 9 through postnatal day (PND) 14, as demonstrated in a series of behavioral and neurochemical studies (Aldridge et al., 2005a; Icenogle et al., 2004; Levin et al., 2001, 2002; Qiao et al., 2003, 2004; Timofeeva et al., 2008b). The first postnatal week of the rat brain development is roughly equivalent to the third trimester of gestation and early postnatal period in humans (Dobbing and Sands, 1979), suggesting that the second and third trimesters of pregnancy as well as postnatal development in humans would also be vulnerable to neurotoxic effects of OPs. Our studies have characterized in rats the neurobehavioral toxicity of three OP pesticides that are widely used in agriculture and for domestic purposes: chlorpyrifos (CPF), diazinon (DZN) and parathion (PTN). Slotkin and colleagues demonstrated that exposure either to 1 or 5 mg/kg/day of CPF, 0.5 or 2 mg/kg/day of DZN, or 0.1 mg/kg/day of PTN, which include the dose ranges subthreshold for overt signs of systemic toxicity, elicit neurotoxic damage of the developing brain (Campbell et al., 1997; Qiao et al., 2003; Roy et al., 2004; Slotkin and Seidler, 2007; Slotkin et al., 2006b, 2008a,b; Song et al., 1997; Whitney et al., 1995). For example, in rats, exposure to the low-dose OPs exerts disruptive effects on neuronal development, with respect to DNA synthesis (Dam et al., 1998), gene transcription (Crumpton et al., 2000), cell differentiation (Roy et al., 1998), and synaptogenesis (Dam et al., 1999a,b). Slotkin and colleagues found that 24 h after daily subcutaneous injections (1 or 5 mg/kg/day of CPF) for 4 days during the postnatal period (PND 1–4), the lower dose of CPF produced only 25% and the higher dose 65% AChE inhibition in the brainstem of postnatal rats (Song et al., 1997). Similar results were reported by Zheng et al. (2000), who found that repeated oral exposure of neonatal rats to 4.5 mg/kg of CPF does not cause any overt signs of toxicity. Betancourt and Carr (Betancourt and Carr, 2004) also reported 23% and 43% inhibition of forebrain AChE in rats after similar oral treatment with 1.5 and 3.5 mg/kg/ day CPF, respectively. The lower and higher doses (0.5 and 2 mg/kg/day) of DZN, injected on PND 1–4 caused 5–10% and 20% of brain AChE inhibition, respectively, 24 h later, and similar treatment with PTN at a dose of 0.1 mg/kg/day caused a statistically significant inhibition that was only

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The effect of early postnatal CPF on learning was sex-selective: male rats increased, while females decreased working memory errors in the 16-arm radial maze (Aldridge et al., 2005a; Levin et al., 2001). Thus, CPF reversed normal sex differences in this function (Figure 33.1). The effect of neonatal DZN and PTN on learning was not sex-selective, but each compound caused different changes in this function. DZN, for example, significantly diminished, and PTN improved learning abilities in the exposed rats. Again, the different outcome could be caused by promotional actions of acetylcholine, because the low dose of PTN caused somewhat greater inhibition of AChE than the low-dose DZN. The developmental OP exposure significantly altered the neural mechanisms used to solve the spatial memory radial-arm maze task. Neonatal rats exposed to DZN demonstrated higher sensitivity to the amnestic effect of the muscarinic receptor antagonist scopolamine, while rats exposed neonatally to PTN and CPF did not. Rats neonatally exposed to CPF demonstrated higher sensitivity to ketanserin, a 5HT2 serotonin receptor antagonist. These results evidence the abnormally increased reliance of cognitive functions, such as learning, on the muscarinic cholinergic system in DZN-exposed rats and on 5HT2 receptors in CPF-treated animals, thus demonstrating that various pesticides differently affect neurotransmitter systems (Aldridge et al., 2005a; Timofeeva et al., 2008a). Unlike neonatal exposure, prenatal exposure to the low doses of CPF, sex-selectively affected sensitivity to the scopolamine challenge with significantly reduced responsivity in females (Icenogle et al., 2004; Levin et al., 2002), indicating a lessened dependence on muscarinic ACh mechanisms for solving the radial maze task.

Early Postnatal Chlorpyrifos Exposure Reverses Normal Sex Differences in Radial-Arm Maze Performance 14 13 Total Errors

5–15% (Slotkin et al., 2006b). These values are far below the 70% threshold necessary for symptoms of cholinergic hyperstimulation (Clegg and van Gemert, 1999). These treatments thus resemble the nonsymptomatic exposures to OP pesticides reported in pregnant women (De Peyster et al., 1993; Souza et al., 2005). Long-term behavioral effects of these low-dose-exposures have been documented. The behavioral studies have focused mainly on three major functional categories: spontaneous activity, emotional response, and cognition. Developmental OP effects on behavioral function have been examined throughout adolescence and into adulthood and compared in both males and females. The battery of tests used in these studies included spontaneous alternation and motor activity in the T-maze, motor activity and habituation in the figure-8 maze, emotional responses in the elevated plus maze, the chocolate milk anhedonia test, the Portsolt forced swim test, the novelty-suppressed feeding test, and learning in the 16-arm radial maze. These studies clearly demonstrated that low-dose short-term exposure of prenatal and neonatal rats to the OP pesticides causes long-lasting neuro­ behavioral impairment. The manifestation and magnitude of such impairments depended on the exposure period, sex, and pesticide used. The diversity of effects between low doses of the different pesticides was particularly noteworthy and supported the contention that non-AChE inhibiting effects of the OP pesticides were important for the persisting neurobehavioral toxicity. Emotional responses and cognitive functions appeared to be more susceptible to the damaging effects of the pesticides than motor activity. For example, recent investigations (Aldridge et al., 2005a; Levin et al., 2001; Roegge et al., 2008; Timofeeva et al., 2008a,b) revealed that neonatal rat exposure to one of the three above-mentioned OP pesticides caused no alterations in motor activity, as tested in the figure-8 maze, and no changes in response latency in the 16-arm radial maze, but did produce significant persistent effects on emotional/cognitive functions. However, it should be noted that some changes in motor activity were observed in rats exposed prenatally to CPF or neonatally to DZN, mostly consisting in altered response to novel environments (Levin et al., 2002; Timofeeva et al., 2008a). Interestingly, animals exposed to the lower but not to the higher doses of one of the three studied pesticides demonstrated changes in learning, thus revealing nonmonotonic dose effect functions of OPs. The nonmonotonic effect, or disappearance of abnormality, with a slightly increased dose of the OP could be a result of positive trophic effect provided by low-level cholinergic stimulation in the developing brain. Acetylcholine serves as a trophic factor that, depending on the developmental stage, promotes neural cell replication and differentiation (Hohmann, 2003; Lauder and Schambra, 1999) and thus at a modest level could offset the damaging effect of the non-AChE inhibitionrelated neurotoxic effects of the OPs.

p<0.01

p<0.025

12 Male

11

Female

10 9 8 Control CPF (1 mg/kg) Postnatal Days 1-4 Treatment

Figure 33.1  Early postnatal chlorpyrifos exposure (1 mg/kg/day, PND 1–4) in rats reverses the normal sex difference in spatial discrimination in the 16-arm radial maze (Aldridge et al., 2005a; Levin et al., 2001).

Chapter | 33  Lasting Behavioral Consequences of Organophosphate Pesticide Exposure During Development

Sex-selective alterations after exposure to OPs were also detected in emotional responses. Thus, males but not females treated in early postnatal period with the lower dose of CPF significantly increased the time spent in the opened arms of the elevated plus maze (Aldridge et al., 2005a). DZN-treated males, on the contrary, decreased their activity in the open arms (Roegge et al., 2008). It was suggested that OP exposure can impair aspects of sexual differentiation of the brain when treatment occurs during the peak period in which these differences are first established (Aldridge et al., 2004). The late gestational through early postnatal period in the rats is considered to be this period (MacLusky and Naftolin, 1981; McCarthy, 1994; Vaccari et al., 1977). Early postnatal exposure to the higher dose of PTN (0.2 mg/kg) also affected the animals’ behavior in the elevated plus maze; however, the changes were not sex-selective: both males and females increased the time spent in the open arms and number of center crossings (Timofeeva et al., 2008b). An increased time spent in the open arms and an increased number of center crossings are thought to reflect reduced anxiety and greater risk taking, the behaviors that are associated with the 5HT system. Slotkin and colleagues have recently shown that neonatal exposure to CPF, DZN, or PTN evoke profound alterations in the 5HT function in rats (Aldridge et al., 2004; Slotkin et al., 2008b,c), but with major differences among the three agents in terms of overall effect, peak period of changes, sex-selectivity, and brain regions affected (Slotkin et al., 2006b). Comparing the ability of the three OPs to elicit immediate and delayed changes in the 5HT systems after early postnatal exposure of rats to doses spanning the threshold of barely detectable cholinesterase inhibition, Slotkin and colleagues found a good correlation between neurochemical and behavioral changes. CPF evoked persistent upregulation of 5HT receptor expression in males (Slotkin and Seidler, 2005), while DZN caused a lasting deficit in 5HT1 receptors, again in males only (Slotkin et al., 2008c), keeping with their differences in emotional behavior evaluated in the elevated plus maze. More recently, they found that comparable exposure to PTN specifically upregulates 5HT1 receptors in the frontal/parietal cortex, peaking by PND 60 (Slotkin et al., 2008b). PND 60 was the age at which increased risk taking behavior in the elevated plus maze was observed in the rats neonatally treated with PTN (Timofeeva et al., 2008b). These findings demonstrate a strong link between neurochemical and emotional alterations caused by different OPs. We also observed that behavioral changes caused by neonatal exposure to low-dose PTN were less widespread and distinctly smaller in magnitude than those elicited by comparable exposures to CPS and DZN. Because neuro­ developmental deficits entail multiple mechanisms unrelated to AChE inhibition, the lower maximum tolerated dose for PTN means that systemic toxicity limits nonsymptomatic exposures to levels substantially lower than those achieved

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with either CPF or DZN. In conjunction with Slotkin and colleagues’ earlier neurochemical findings (Slotkin et al., 2006a), this reinforces the complete dichotomy between systemic toxicity of OPs and their propensity to elicit developmental neurotoxicity. OPs are usually thought of as neurotoxicants. However, recent investigations made in the same laboratory (Lassiter et al., 2008a,b) revealed that OPs have other targets that contribute to morbidity, including metabolic disturbances that have a potential impact on obesity and diabetes. This discovery opens a new avenue for laboratory and epidemiological investigations and emphasizes the diversity of OPs’ detrimental effects on health. Thus, studies conducted on rats models demonstrated that pre- and postnatal exposure to OP pesticides, such as CPF, DZN, and PTN, at doses that elicit barely detectable cholinesterase inhibition and span the signs of systemic toxicity, nevertheless produce behavioral alterations lasting into adolescence and adulthood. These studies support the idea that, in the developing brain, various OPs target specific neurotransmitter systems differently from each other, involving mechanisms beyond their shared property of cholinesterase inhibition. Strong scientific evidence that low doses of OPs cause developmental neurotoxicity in rats supports the idea that exposure to low doses of OP pesticides is detrimental for children’s health and development.

33.4  Mouse models Behavioral alterations have been also extensively documented in mice species following similar low-level gestational and postnatal exposures to CPF, which do not substantially inhibit brain AChE (Laviola et al., 2006; Ricceri et al., 2003, 2006; Venerosi et al., 2008). For example, a modest AChE inhibition (about 25%) was found in mice (an outbred Swiss-derived CD-1 strain) treated with CPF (3 mg/kg/day, sc) only on PND 1–4, but not after treatment on PND 11–14, 1 h but not 24 h after the last CPF administration. The percentage of inhibition reported for these mice was even lower than that reported for rats of the same age and for similar CPF doses and testing time. Moreover, the same authors demonstrated that AChE activity recovered more quickly in mice. Such differences likely result from species-specific differences. Nevertheless, CPF exposure elicited multiple behavioral alterations in adolescent and adult mice, including motor activity as well as emotional and social responses. Thus, mouse studies also contribute strong evidence to the concept that OP pesticides affect developing brain via mechanisms different from AChE inhibition. After exposing mice (PND 1–4) to 1 or 3 mg/kg/day CPF, Ricceri and colleagues (2003) investigated their behavior at different ages: ultrasonic vocalization and homing at neonatal stage (PND 5, 8, 11), locomotor activity at weaning (PND 25), novelty seeking in adolescence,

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social interactions in young adulthood (PND 45), and passive avoidance learning at PND 60. As with rats, some of the tested behaviors were affected by the exposure and some were not, emphasizing the fact that CPF targeted only particular functions and neurotransmitter systems and does not affect brain systems globally. Similar to rats (Icenogle et al., 2004; Levin et al., 2001; Timofeeva et al., 2008a), a CPF-induced increase in locomotor activity in mice was detected in novel environments (Ricceri et al., 2003) and appears to reflect emotional response. Low-dose CPF also significantly affected social behavior, which has not to date been tested in rats. The social behavior repertoire analyzed on PND 45 indicated that CPF at the higher dose of 3 mg/kg/day increased the expression of soliciting responses in both postnatal groups and in both sexes. Such behavior suggests either higher propensity to interact with a social mate or enhanced arousal in responding to social cues. When the authors evaluated aggressive behavior as a whole, they found that the lower dose of CPF was most effective in producing more agonistic responses in males, reflecting a nonmonotonic and sex-selective effect of the OP. Similar sex-selective nonmonotonic behavioral responses to CPF have been seen with regard to cognitive behavior in rats. Ricceri and colleagues suggested that the presence of aggressive components in male social behavior at a stage when the affiliative/social components of interaction should be prevalent might express an alteration in the normal development of social behavioral patterns. The same group also investigated behavioral effects in adult mice after either fetal (GD 15–18) and/or neonatal (PND 11–14) CPF exposure at doses near threshold for inhibiting fetal and neonatal brain cholinesterase (Ricceri et al., 2006). Although serum AChE activity was slightly inhibited 24 h after the last postnatal treatment, no changes in brain AChE activity were detected later in the study. Nevertheless, selective behavioral impairments were observed in the adult mice. As in previous studies, exposed mice showed increased locomotion in the open field, and enhanced aggressive response in the socioagonistic behavior after postnatal CPF. This study presented a new finding, further demonstrating that locomotion in a novel environment and social behavior are sensitive functions to the impairing effect of developmental CPF. Moreover, postnatal CPF also increased maternal responsiveness toward pups by virgin mice. Both socioagonistic and maternal behaviors are sex dimorphic, their performance depends on the maturation of sex hormone-regulated brain pathways. Thus, there is a potential for endocrine-disrupting activity of OPs. Prenatal treatment was less deleterious: it increased only offensive posture frequency. Postnatal CPF also reduced anxiety in the elevated plus maze in both sexes, with females more severely affected. Reduced anxiety has been observed in adult male rats exposed to CPF and in both sexes exposed to PTN postnatally (Aldridge et al., 2005a; Timofeeva et al., 2008b). Interestingly, in mice, CPF affected preferentially females’ behavior, raising

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the open arm time to the same levels found in males. There are strong species-specific differences between mice and rats, as female mice, in contrast to rats, are more anxious than males when confronted with a novel environment (Augustsson et al., 2005). Interestingly, the authors reported that they did not find CPF-induced impairment in water-maze acquisition. In the following study (Venerosi et al., 2008), they found that exposure to CPF (PND 11–14, 3 mg/kg /day, sc) did not interfere with social behavior and social preferences in adolescence, whereas in adult lactating mice it induced significant behavioral alterations in different aspects of the maternal repertoire. Motivation to build and defend their nest was significantly decreased in CPF-treated females, which were also less anxious than controls in the dark-light paradigm. CPF-treated females tended to attack the male intruder less than control females and showed a significant decrease of defensive responses when attacked. In addition, investigative and subordinate behaviors toward the intruder were significantly enhanced, apparently at the expense of maternal aggressive patterns. The authors proposed a hypothesis that CPF exposure in a critical neonatal phase might affect the development of the hypothalamic neuroendocrine circuits regulating maternal social responses. A recent study by Laviola et al. (2006) investigated the interplay between genetic vulnerability and prenatal (GD 14–16) exposure to a neurotoxic compound such as CPF-oxon (0.144 mg/day, corresponding to approximately 5 mg/kg/day of CPF, administered with osmotic minipumps). The “reeler” mice, lacking the extracellular-matrix protein reelin, were used. They found complex interactions between genetic (reeler genotype) and epigenetic (prenatal exposure to CPF-oxon) factors. CPF-oxon paradoxically reversed the effects produced by reelin absence in such parameters as ultrasound vocalization (PND 7), amphetamine-induced locomotion, and stereotypy (PND  70). CPF-oxon also accelerated the maturation of grasping reflex in controls, when measured in pups at PND 3, the effects that were counteracted by progressive reelin absence. Finally, for parameters such as righting reflex latency and scopolamine-induced locomotor activity, prenatal CPF-oxon unmasked the otherwise latent genotype deficiency. The effects observed in this study cannot be explained by enduring AChE blockade, since CPF-oxon induced AChE blockade is reversible in a few hours (Betancourt and Carr, 2004) in rats and AChE activity recovers even quicker in mice than in rats (Ricceri et al., 2003, 2006). This study offers a new avenue of investigation, the possibility of unmasking genetic deficiency with environmental factors such as OP pesticides. This approach could help model the developmental pathogenesis of neurologic diseases, such as autism, Parkinson disease, and others. Thus, mouse studies together with rat models evidence that brief developmental exposure to OP pesticides at doses

Chapter | 33  Lasting Behavioral Consequences of Organophosphate Pesticide Exposure During Development

Zebrafish models have emerged as a promising sensitive tool enabling quick and reliable detection of neurodevelopmental injuries in response to OP pesticide exposure (Levin et al., 2003, 2004; Linney et al., 2004). Zebrafish with their clear chorion and extensive developmental information base, also provide a valuable model for assessment of molecular processes underlying neurodevelopmental impairments. Levin et al. (2003) developed methods for assessing spatial discrimination learning in zebra­fish, which can differentiate response latency from choice accuracy in a three-chambered fish tank. Low and high doses of CPF (10 and 100 ng/ml), administered to zebra­fish embryos for the first 5 days postfertilization, both had significant persistent effects on spatial discrimination and response latency over 18 weeks of testing (Figure 33.2). The authors found that the high but not the low dose significantly accelerated the mortality rate in 20- to 38-week-old fish when the study was done. The lower-dose effect was observed mainly in early testing, while impairment caused by the higher dose became more pronounced with continued testing. The higher dose also caused more pervasive impairment. The two doses had opposite effects on response latency, with the low dose significantly increasing and the high dose decreasing response latency. In another study, Levin et al. (2004) studied the effect of the same CPF treatment on zebrafish swimming activity. They found that the higher dose produced a significant slowing of swimming activity on days 6–9 postfertilization and had a persistent effect of impairing spatial discrimination and decreasing response latency into adulthood. More recently, developmental exposure of zebrafish to CPF (first 5 days postfertilization to 100 ng/ml or 0.29 mol) has been shown to cause persistent deficits in brain dopamine levels and hyperactive response to startling tactile stimuli (Eddins et al., 2010). These studies clearly demonstrate that developmental CPF caused behavioral alterations in zebrafish, lasting throughout adulthood. The zebrafish models could serve as quick and inexpensive tools for screening OP pesticides on their toxicity. Animal studies have therefore demonstrated a strong link between fetal/postnatal exposure to a specific OP pesticide, its dose, and long-lasting behavioral alterations, including emotional and social responses, cognition, and spontaneous activity. These results justify public health concerns

70 65 Percent Correct

33.5  Fish models

Developmental Chlorpyrifos Exposure Effects on Average Choice Accuracy

60

* **

55 50

vs. Control * p<0.05 ** p<0.01

45 40 (A)

0

10 100 Chlorpyrifos (ng/ml)

Developmental Chlorpyrifos Exposure Effects on Average Response Latency 30 **

25 Seconds per Trial

barely inhibiting, if at all, brain AChE activity, induce long-term behavioral effects, including altered spontaneous activity, emotional response, learning and memory, and social interaction. Hence, animal data not only justify our concern for the safety of children living in environments contaminated with OPs, but also show the mechanisms of the neurodevelopmental impairments and its prognosis.

843

20

**

15 10 5

vs. Control ** p<0.01

0

(B)

0

10 100 Chlorpyrifos (ng/ml)

Figure 33.2  In the three-chamber delayed spatial alternation memory test for zebrafish (A) memory accuracy was significantly reduced in adult zebrafish after early developmental exposure to 10 or 100 ng/ml of chlorpyrifos for the first 5 days post-fertilization. (B) There was a biphasic effect on swimming activity with the 10 ng/ml dose causing hypoactivity, and the 100 ng/ml causing hyperactivity (Levin et al., 2003)

about the safety of children living in the contaminated environment and could be useful in developing treatment and protective strategy.

Conclusions Experimental studies in rodent and piscine species have clearly demonstrated the vulnerability of the developing nervous system to OP pesticide exposure. This is clearly expressed in persistent behavioral disruption. OPs affect emotional and cognitive functions, social responses, and sex-related behavioral patterns, and in addition disrupt mechanisms of weight regulation. Significantly altered behavioral function was still seen long after the short-term, low-dose exposure to a variety of OPs. The fact that the

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character of the neurobehavioral dysfunction varied across the different OPs at doses with little or no AChE inhibition implies that AChE inhibition is not a good toxicodynamic metric by which to judge the potential toxic risk of OP pesticides. Other mechanisms of neurotoxicity should be considered. Experimental animal studies support the epidemiological findings that OP pesticides represent a serious danger to neurobehavioral development.

Acknowledgments This research was supported by Duke University Superfund Basic Research Center (NIH ES10356).

References Aldridge, J. E., Seidler, F. J., Meyer, A., Thillai, I., and Slotkin, T. A. (2003). Serotonergic systems targeted by developmental exposure to chlorpyrifos: effects during different critical periods. Environ. Health Perspect. 111, 1736–1743. Aldridge, J. E., Seidler, F. J., and Slotkin, T. A. (2004). Developmental exposure to chlorpyrifos elicits sex-selective alterations of serotonergic synaptic function in adulthood: critical periods and regional selectivity for effects on the serotonin transporter, receptor subtypes, and cell signaling. Environ. Health Perspect. 112, 148–155. Aldridge, J. E., Levin, E. D., Seidler, F. J., and Slotkin, T. A. (2005a). Developmental exposure of rats to chlorpyrifos leads to behavioral alterations in adulthood, involving serotonergic mechanisms and resembling animal models of depression. Environ. Health Perspect. 113, 527–531. Aldridge, J. E., Meyer, A., Seidler, F. J., and Slotkin, T. A. (2005b). Alterations in central nervous system serotonergic and dopaminergic synaptic activity in adulthood after prenatal or neonatal chlorpyrifos exposure. Environ. Health Perspect. 113, 1027–1031. Ascherio, A., Chen, H., Weisskopf, M. G., O’Reilly, E., McCullough, M. L., Calle, E. E., Schwarzschild, M. A., and Thun, M. J. (2006). Pesticide exposure and risk for Parkinson’s disease. Ann. Neurol. 60, 197–203. Augustsson, H., Dahlborn, K., and Meyerson, B. J. (2005). Exploration and risk assessment in female wild house mice (Mus musculus musculus) and two laboratory strains. Physiol. Behav. 84, 265–277. Barbeau, A., Roy, M., Bernier, G., Campanella, G., and Paris, S. (1987). Ecogenetics of Parkinson’s disease: prevalence and environmental aspects in rural areas. Can. J. Neurol. Sci. 14, 36–41. Berkowitz, G. S., Wetmur, J. G., Birman-Deych, E., Obel, J., Lapinski, R. H., Godbold, J. H., Holzman, I. R., and Wolff, M. S. (2004). In utero pesticide exposure, maternal paraoxonase activity, and head circumference. Environ. Health Perspect. 112, 388–391. Betancourt, A. M., and Carr, R. L. (2004). The effect of chlorpyrifos and chlorpyrifos-oxon on brain cholinesterase, muscarinic receptor binding, and neurotrophin levels in rats following early postnatal exposure. Toxicol. Sci. 77, 63–71. Brent, R. L., and Weitzman, M. (2004). The current state of knowledge about the effects, risks, and science of children’s environmental exposures. Pediatrics 113, 1158–1166. Campbell, C. G., Seidler, F. J., and Slotkin, T. A. (1997). Chlorpyrifos interferes with cell development in rat brain regions. Brain. Res. Bull. 43, 179–189.

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Clegg, D. J., and van Gemert, M. (1999). Determination of the reference dose for chlorpyrifos: proceedings of an expert panel. J. Toxicol. Environ. Health B. Crit. Rev. 2, 211–255. Colborn, T. (2006). A case for revisiting the safety of pesticides: a closer look at neurodevelopment. Environ. Health Perspect. 114, 10–17. Colosio, C., Tiramani, M., and Maroni, M. (2003). Neurobehavioral effects of pesticides: state of the art. Neurotoxicology 24, 577–591. Crumpton, T. L., Seidler, F. J., and Slotkin, T. A. (2000). Developmental neurotoxicity of chlorpyrifos in vivo and in vitro: effects on nuclear transcription factors involved in cell replication and differentiation. Brain Res. 857, 87–98. Dam, K., Seidler, F. J., and Slotkin, T. A. (1998). Developmental neuro­ toxicity of chlorpyrifos: delayed targeting of DNA synthesis after repeated administration. Brain Res. Dev. Brain Res. 108, 39–45. Dam, K., Garcia, S. J., Seidler, F. J., and Slotkin, T. A. (1999a). Neonatal chlorpyrifos exposure alters synaptic development and neuronal activity in cholinergic and catecholaminergic pathways. Brain Res. Dev. Brain Res. 116, 9–20. Dam, K., Seidler, F. J., and Slotkin, T. A. (1999b). Chlorpyrifos releases norepinephrine from adult and neonatal rat brain synaptosomes. Brain Res. Dev. Brain Res. 118, 129–133. Daniell, W., Barnhart, S., Demers, P., Costa, L. G., Eaton, D. L., Miller, M., and Rosenstock, L. (1992). Neuropsychological performance among agricultural pesticide applicators. Environ. Res. 59, 217–228. De Peyster, A., Willis, W. O., Molgaard, C. A., MacKendrick, T. M., and Walker, C. (1993). Cholinesterase and self-reported pesticide exposure among pregnant women. Arch. Environ. Health 48, 348–352. Dobbing, J., and Sands, J. (1979). Comparative aspects of the brain growth spurt. Early Hum. Dev. 3, 79–83. Eddins, D., Cerutti, D., Williams, P., Linney, E., and Levin, E. D. (2010). Developmental chlorpyrifos causes behavioral and neurochemical defects in zebrafish. Neurotoxicol. Teratol. in press. Eskenazi, B., Marks, A. R., Bradman, A., Harley, K., Barr, D. B., Johnson, C., Morga, N., and Jewell, N. P. (2007). Organophosphate pesticide exposure and neurodevelopment in young Mexican-American children. Environ. Health Perspect. 115, 792–798. Faustman, E. M., Silbernagel, S. M., Fenske, R. A., Burbacher, T. M., and Ponce, R. A. (2000). Mechanisms underlying children’s susceptibility to environmental toxicants. Environ. Health Perspect. 108 (Suppl. 1), 13–21. Fiedler, N., Kipen, H., Kelly-McNeil, K., and Fenske, R. (1997). Longterm use of organophosphates and neuropsychological performance. Am. J. Ind. Med. 32, 487–496. Garcia, S. J., Seidler, F. J., Crumpton, T. L., and Slotkin, T. A. (2001). Does the developmental neurotoxicity of chlorpyrifos involve glial targets? Macromolecule synthesis, adenylyl cyclase signaling, nuclear transcription factors, and formation of reactive oxygen in C6 glioma cells. Brain Res. 891, 54–68. Garry, V. F., Harkins, M. E., Erickson, L. L., Long-Simpson, L. K., Holland, S. E., and Burroughs, B. L. (2002). Birth defects, season of conception, and sex of children born to pesticide applicators living in the Red River Valley of Minnesota, USA. Environ. Health Perspect. 110(Suppl. 3), 441–449. Hohmann, C. F. (2003). A morphogenetic role for acetylcholine in mouse cerebral neocortex. Neurosci. Biobehav. Rev. 27, 351–363. Icenogle, L. M., Christopher, N. C., Blackwelder, W. P., Caldwell, D. P., Qiao, D., Seidler, F. J., Slotkin, T. A., and Levin, E. D. (2004). Behavioral alterations in adolescent and adult rats caused by a brief subtoxic exposure to chlorpyrifos during neurulation. Neurotoxicol. Teratol. 26, 95–101.

Chapter | 33  Lasting Behavioral Consequences of Organophosphate Pesticide Exposure During Development

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Qiao, D., Seidler, F. J., Tate, C. A., Cousins, M. M., and Slotkin, T. A. (2003). Fetal chlorpyrifos exposure: adverse effects on brain cell development and cholinergic biomarkers emerge postnatally and continue into adolescence and adulthood. Environ. Health Perspect. 111, 536–544. Qiao, D., Seidler, F. J., Abreu-Villaca, Y., Tate, C. A., Cousins, M. M., and Slotkin, T. A. (2004). Chlorpyrifos exposure during neurulation: cholinergic synaptic dysfunction and cellular alterations in brain regions at adolescence and adulthood. Brain Res. Dev. Brain Res. 148, 43–52. Quandt, S. A., Arcury, T. A., Rao, P., Snively, B. M., Camann, D. E., Doran, A. M., Yau, A. Y., Hoppin, J. A., and Jackson, D. S. (2004). Agricultural and residential pesticides in wipe samples from farmworker family residences in North Carolina and Virginia. Environ. Health Perspect. 112, 382–387. Raugh, V. A., Garfinkel, R., Perera, F. P., Andrews, H. F., Hoepner, L., Barr, D. B., Whitehead, R., Tang, D., and Whyatt, R. W. (2006). Impact of prenatal chlorpyrifos exposure on neurodevelopment in the first 3 years of life among inner-city children. Pediatrics 118, e1845–e1859. Ray, D. E., and Richards, P. G. (2001). The potential for toxic effects of chronic, low-dose exposure to organophosphates. Toxicol. Lett. 120, 343–351. Ricceri, L., Markina, N., Valanzano, A., Fortuna, S., Cometa, M. F., Meneguz, A., and Calamandrei, G. (2003). Developmental exposure to chlorpyrifos alters reactivity to environmental and social cues in adolescent mice. Toxicol. Appl. Pharmacol. 191, 189–201. Ricceri, L., Venerosi, A., Capone, F., Cometa, M. F., Lorenzini, P., Fortuna, S., and Calamandrei, G. (2006). Developmental neurotoxicity of organophosphorous pesticides: fetal and neonatal exposure to chlorpyrifos alters sex-specific behaviors at adulthood in mice. Toxicol. Sci. 93, 105–113. Ritz, B., and Yu, F. (2000). Parkinson’s disease mortality and pesticide exposure in California 1984-1994. Int. J. Epidemiol. 29, 323–329. Roegge, C. S., Timofeeva, O. A., Seidler, F. J., Slotkin, T. A., and Levin, E. D. (2008). Developmental diazinon neurotoxicity in rats: later effects on emotional response. Brain Res. Bull. 75, 166–172. Roy, T. S., Andrews, J. E., Seidler, F. J., and Slotkin, T. A. (1998). Chlorpyrifos elicits mitotic abnormalities and apoptosis in neuroepithelium of cultured rat embryos. Teratology 58, 62–68. Roy, T. S., Seidler, F. J., and Slotkin, T. A. (2004). Morphologic effects of subtoxic neonatal chlorpyrifos exposure in developing rat brain: regionally selective alterations in neurons and glia. Brain Res. Dev. Brain Res. 148, 197–206. Slotkin, T. A., and Seidler, F. J. (2005). The alterations in CNS serotonergic mechanisms caused by neonatal chlorpyrifos exposure are permanent. Brain Res. Dev. Brain Res. 158, 115–119. Slotkin, T. A., and Seidler, F. J. (2007). Comparative developmental neurotoxicity of organophosphates in vivo: transcriptional responses of pathways for brain cell development, cell signaling, cytotoxicity and neurotransmitter systems. Brain Res. Bull. 72, 232–274. Slotkin, T. A., Levin, E. D., and Seidler, F. J. (2006a). Comparative developmental neurotoxicity of organophosphate insecticides: effects on brain development are separable from systemic toxicity. Environ. Health Perspect. 114, 746–751. Slotkin, T. A., Tate, C. A., Ryde, I. T., Levin, E. D., and Seidler, F. J. (2006b). Organophosphate insecticides target the serotonergic system in developing rat brain regions: disparate effects of diazinon and parathion at doses spanning the threshold for cholinesterase inhibition. Environ. Health Perspect. 114, 1542–1546.

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Slotkin, T. A., Bodwell, B. E., Levin, E. D., and Seidler, F. J. (2008a). Neonatal exposure to low doses of diazinon: long-term effects on neural cell development and acetylcholine systems. Environ. Health Perspect. 116, 340–348. Slotkin, T. A., Levin, E. D., and Seidler, F. J. (2008b). Developmental neuro­toxicity of parathion: Progressive effects on serotonergic systems in adolescence and adulthood. Neurotoxicol. Teratol. 75, 640–647. Slotkin, T. A., Ryde, I. T., Levin, E. D., and Seidler, F. J. (2008c). Developmental neurotoxicity of low dose diazinon exposure of neonatal rats: effects on serotonin systems in adolescence and adulthood. Brain Res. Bull. 75, 640–647. Song, X., Seidler, F. J., Saleh, J. L., Zhang, J., Padilla, S., and Slotkin, T. A. (1997). Cellular mechanisms for developmental toxicity of chlorpyrifos: targeting the adenylyl cyclase signaling cascade. Toxicol. Appl. Pharmacol. 145, 158–174. Souza, M. S., Magnarelli, G. G., Rovedatti, M. G., Cruz, S. S., and De D’Angelo, A. M. (2005). Prenatal exposure to pesticides: analysis of human placental acetylcholinesterase, glutathione S-transferase and catalase as biomarkers of effect. Biomarkers 10, 376–389. Tilson, H. A. (2000). Neurotoxicology risk assessment guidelines: developmental neurotoxicology. Neurotoxicology 21, 189–194. Timofeeva, O. A., Roegge, C. S., Seidler, F. J., Slotkin, T. A., and Levin, E. D. (2008a). Persistent cognitive alterations in rats after early postnatal exposure to low doses of the organophosphate pesticide, diazinon. Neurotoxicol. Teratol. 30, 38–45. Timofeeva, O. A., Sanders, D., Seemann, K., Yang, L., Hermanson, D., Regenbogen, S., Agoos, S., Kallepalli, A., Rastogi, A., Braddy, D., Wells, C., Perraut, C., Seidler, F. J., Slotkin, T. A., and Levin, E. D. (2008b). Persistent behavioral alterations in rats neonatally exposed

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Chapter 34

The Nonhuman Primate as a Translational Model for Pesticide Research Merle G. Paule, Ph.D. National Center for Toxicological Research, Food and Drug Administration, Jefferson, Arkansas

34.1  Introduction Our need to use animal models in risk-assessment and riskbenefit analyses derives directly from our desire to know, a priori, what is going to happen to us humans when we are exposed to a given chemical or procedure. Unfortunately, even after extensive research using animal models, it is often the case that adverse effects are seen in humans that were not predicted by the data obtained from our animal models; this is perhaps most obvious in the case of therapeutic agents that fail in clinical trials after extensive preclinical (animal) research. This means that either the animal models used to generate such data were inappropriate or the studies that employed the animal model(s) were not designed properly, or both. In the case of pesticides, others have argued for revisiting their safety primarily because of the vast numbers of pesticides on the market and the large number of possible target tissues and endpoints that can be affected and that can differ depending upon the timing and life stage at which exposure occurs (Colborn, 2006). Unfortunately, there are ample data on the toxicities associated with high dose, acute exposures to pesticides in humans, typically through accidents or occupational exposures (e.g., Lotti, 2001; Kamel et al., 2005) and organo­ phosphate induced delayed neuropathy (OPIDN) is a (currently the only) known human neurodegenerative disease associated with pesticide exposure (Doherty, 2006). While there have been recent suggestions that pesticide exposures may be risk factors for Parkinson’s disease, the evidence in support of this is inconclusive and existing guidelines do not specify protocols that would address this possibility, highlighting the need for additional animal models and studies (Doherty, 2006). There is little evidence that moderate pesticide exposures result in adverse effects, perhaps because fewer studies have examined this issue (Kamel and Hoppin, 2004; Kamel et al., 2005). And, perhaps most importantly, the data needed to determine Hayes’ Handbook of Pesticide Toxicology

the consequences of chronic exposure to low doses of pesticides—scenarios experienced by the bulk of the population—are indeed limited (Alavanja et al., 2004) but suggest a relationship between urinary organophosphate metabolite levels and acetylcholinesterase inhibition (i.e., pesticide exposures) and deficits in neurobehavioral performance (Rothlein et al., 2006) and neuropsychiatric, motor, and extrapyramidal symptoms (Salvi et al., 2003), respectively. It is also clear that quantitation of acetylcholinesterase inhibition in the face of OP exposure is insufficient for understanding potentially related neurotoxicity (Salvi et al., 2003). Numerous authors have also highlighted the need for better models and more studies that better reflect likely human contact with pesticides (e.g., Doherty, 2006; Kamel and Hoppin, 2004; Seegal, 1996). Recent findings on birth outcomes suggest that pesticide exposures are related to decreased birth weights and length of pregnancy (Perera et al, 2005). The case has also been made, albeit for polychlorinated biphenyls (PCBs), that human epidemiological data are not likely to allow for the determination of the adverse effects of a particular chemical because of the background of additional contaminants in the populations being studied. Thus, it has been suggested that future risk assessments should rely more heavily on laboratoryderived data, including studies in nonhuman primates (Seegal, 1996). The same argument is valid for pesticide exposures.

34.2  Animal model considerations Since the majority of human encounters with pesticides are likely of the low dose, chronic/lifetime variety, it seems obvious that those kinds of exposure scenarios should serve as the blueprint after which animal studies should be modeled. In addition, assessment of actual pesticide exposure in humans needs to be improved (Kamel and 847

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Hoppin, 2004) so that exposure levels in the population are known more precisely and can be appropriately configured in our animal exposure models. While it is not the purpose of this chapter to review the ubiquitous nature of human pesticide exposure, it is noteworthy that inner-city children appear to be at a relatively higher risk of such exposure and that children represent a population highly vulnerable to pesticides (Landrigan et al., 1999). Entire volumes have been devoted to the discussion of pesticides in the diets of children (e.g., National Research Council Staff, 1993) and in which the special characteristics of children have been addressed (e.g., Chapter 2 of that publication), as has the need for a better pediatric model. Some of these unique characteristics of children stem not only from their proximity to the ground, high hand-to-mouth behaviors, dietary patterns, and propensity to absorb more pesticides from their environment than adults (Landrigan et al., 1999), but also from the fact that insults experienced during development can have long-lasting, perhaps lifetime consequences (Weiss, 2000). Thus, in general, there seems to be a growing consensus that the past and current animal models and approaches used to assess pesticide toxicity have not and are not providing the kinds of data that are likely to be of relevance to the most typical human exposures. It seems that there is a clear and growing need to consider alternative approaches. Again, given that typical interactions of the majority of the population with pesticides involve chronic, low-dose exposures, high priority should be afforded those laboratory animal studies concerned with the effects of chronic exposures to low doses. Given that children represent a portion of the population likely to be most at risk for pesticide exposure and effect, our animal models must contain, or focus on, a pediatric equivalent. Perhaps a second sensitive population should include aged animals to serve as surrogates for persons with declining physiological and cognitive reserves. While it appears that much of what is known about the toxicology of pesticides has been obtained from relatively high exposures in normal adult animals and humans, the adult human population may be the least likely to be affected by typical pesticide exposures and, thus, may be the least relevant population around which to build our models. Given also that most of the toxicities associated with pesticide exposures are referable to the nervous system, resources should be focused foremost on determining the effects of pesticides on the development, function, and integrity of that system.

34.2.1  Types of Animal Models The history and use of nonhuman animals as models of humans has been discussed by others in some detail and will not be belabored here. Basically, four types of animal models have been described and these include: induced,

Hayes’ Handbook of Pesticide Toxicology

spontaneous, negative, and orphan (Salen, 1994). Interest­ ingly, all of these types of animal models relate to their use in the study of the causes, nature, and cures or treatments of human disorders, rather than to their use as surrogates for normal humans undergoing toxic exposure. Induced models are those in which some abnormal con­dition or disease has been created such as MPTP- or 6-hydroxydopamine-induced Parkinson’s disease (Petzinger and Jakowec, 2005; Soderstrom et al., 2006) or streptozotocininduced diabetes (Akbarzadeh et al., 2007). Recently, and to a growing degree, transgenic techniques are being used to induce specific types of traits in animals: here foreign DNA is artificially inserted into the model’s genome in order to impart to it known characteristics of the model/ disorder of interest. Spontaneous models are animals in which the disorder of interest is expressed naturally, such as in the case of the genetically epileptic baboon, Papio papio (Naquet 1972; Naquet et al., 1995) and the narcoleptic mini-horse (Lunn et al., 1993). Negative animal models are those in which the disorder of interest cannot be observed or induced, thereby making them valuable from the standpoint of providing some understanding of the mechanisms underlying their resistance. An example would be the inability to establish gonococcal infections in rabbits and other resistant species (Ngampasutadol et al., 2008). Orphan models are animal models that have a condition that occurs naturally in them but that has not yet been observed in humans. Again, an interesting point with all of these models is the focus on some sort of derangement from normal that will help in the study of correcting similar derangements in humans or other animals. In the drug development process, the rate-limiting step in the progress of developing a treatment or cure may often be the preclinical model being used to screen for and identify new therapeutic agents and this is generally an animal model or some subsystem(s) derived from animals. The more relevant or appropriate the animal model, i.e., the more it models the human condition for which it is being used, the more likely it will prove useful in development of useful drugs. The importance here again is that the target of the effort is to use animal models to discover chemicals that will cause an effect in humans and, hopefully, that effect will be to normalize function in humans suffering from disease or injury.

34.3  Considerations for animal models of pesticide exposures Pesticide research, on the other hand, requires animal models that should most closely approximate normal human beings so that they can provide important information concerning the likelihood of an adverse outcome in exposed persons. In this manner, the animal model needs to serve as

Chapter | 34  The Nonhuman Primate as a Translational Model for Pesticide Research

a surrogate for healthy humans and, thus, it is of some importance to identify the most appropriate animal for the job. Prime considerations should include: physiology (Is the physiology of the animal model similar or relevant to that of humans?); metabolism (Does the animal model metabolize the pesticide of interest in a fashion similar to that observed in humans? Does the animal exhibit metabolic profiles that are similar to those of humans?); endocrinology (Does the animal model exhibit hormonal patterns that approximate those of humans? Does the animal model posses endocrine and reproductive systems that are similar or relevant to those seen in humans?); placentation/reproductive organs (Does the animal model exhibit reproductive tissue/systems that are similar or relevant to humans?); pharmacology (Does the animal model respond to drugs in a fashion that resembles human responses? Such similarity can tell us something about the validity of the model and the likelihood that it will produce data predictive of other chemical effects in humans); developmental stages and time-course (Does the animal model exhibit all of the life stages seen in humans and do they occur over an appropriate time course?); validity of endpoints (Do the endpoints monitored in our animal models have validity/ relevance to humans?). In general, a continuum of similarities exists between animals and humans with fewer similarities occurring the farther one goes down the phylogenetic tree. It should be noted, however, that phylogenetic proximity does not guarantee predictability. Take, for example, the observation that chimpanzees, arguably our closest relatives phylogenetically, are poor models for AIDS research (King, 1986). Despite decades of experience, toxicologists often choose their animal models based on historical precedent, convenience, cost, or familiarity rather than their appropriateness to model humans (Svendsen, 1994). A primary aim of this chapter is to provide arguments that nonhuman primates are the most appropriate animal models for the conduct of research aimed at determining the adverse effects of pesticide exposures under conditions most likely to be encountered by humans. A discussion of the ethical issues concerning the use of these animals in research can be found in Evans (1990). The current popularity of antianimal research sentiment should not be allowed to deter the pursuit of sound science and the identification and use of the best animal models for use in answering the important human health issues at hand.

34.3.1  Doses, Timing, and Duration of Exposures and Endpoint Considerations Given the relatively ubiquitous nature of pesticides in the environment, exposure, albeit to low doses, likely begins at the moment of conception and continues throughout the

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lifespan. Thus, a critical step in the process of model development would be to adopt this exposure model, understanding that critical periods of development are relatively well defined for a variety of animal models and humans (Rice and Barone, 2000) and important similarities can be identified and exploited. Next, it will be important to identify or define the endpoints of most relevance for humans. As mentioned earlier, in the case of human exposures to pesticides, it seems clear that the nervous system is the system of greatest concern. This is also true for dioxinlike compounds and researchers in this area have called for the development and standardization of neurobehavioral test batteries that can serve as reliable measures of disease endpoints (Lindstrom et al., 1995). Excellent examples of the utility of this approach can be found in studies by Schantz et al. (1991) and Rice (1999) in which monkeys were exposed chronically to relevant doses of polychlorinated biphenyls (PCBs) either during pregnancy and lactation or from birth to 20 weeks of age, respectively. PCB exposures resulted in deficits in spatial learning and memory and reversal tasks, retarded learning, perseverative behavior, and a decrement in the ability to inhibit inappropriate behaviors. Interestingly, these findings in nonhuman primates support similar reports of adverse PCB effects in human children (Winneke et al., 1998). In addition, similar approaches have been applied to studies on the cognitive effects of exposure to endocrine-disrupting chemicals, in general (Schantz and Widholm, 2001). Identifying relevant and translatable biomarkers of nervous system integrity is key for designing appropriate studies with which to explore the adverse events associated with pesticide exposures. Such an accomplishment will also serve the broader neuroscience community at large by helping to harmonize standard models of cognitive function and impairment (e.g., Hachinski et al., 2006). Here again, it will be the purpose of this chapter to describe why the nonhuman primate has demonstrated the best fit.

34.3.2  The Young as Sensitve Targets Given that the nervous systems of the young and the elderly are likely to be the most vulnerable to xenobiotic insult, these populations should receive special attention, with most emphasis being afforded the developing nervous system since derangements occurring early during maturation may have permanent, thus life-long, consequences (Weiss, 2000). Ideally, one would want to repeatedly assess the same subjects for extended periods of time, perhaps even throughout the entire lifespan. It is hard to envision how that would be possible using invasive procedures, thus the utilization of noninvasive or minimally invasive procedures for obtaining metrics of nervous system integrity will be a major focus. These procedures can include

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b­ iochemical, electrophysiological, behavioral, and visual (imaging) approaches.

34.3.3  Endpoints of Toxicity 34.3.3.1  Biochemistry and the “Omics” Interpretation of biochemical and electrophysiological changes can, however, be problematic in terms of translating animal observations to human outcomes. For example, the “omics” technologies are very capable of producing large quantities of biochemical data concerning changes in gene expression (genomics), protein synthesis (proteomics), and metabolic pathways (metabolomics) and expert panels are currently attempting to formulate plans for a scientific infrastructure that will help deal with the very large amounts of data that can be generated using these approaches (National Institutes of Health, 2008). Not surprisingly, a tremendous amount of data gathering, analyses, and validation will be necessary to further our comprehension of the utility of each of the omics approaches and how best to take advantage of them. When, for example, is (are) the best time(s) after an exposure to collect omics data? How do the omic endpoints vary with time of day, month, year, sex, tissue type, and age, and are they different after acute exposures versus chronic exposures? Thus, while the omics technologies hold the promise of the future and even now are providing important information concerning the underlying mechanisms associated with a variety of toxicities, it will likely be some time before they can be harnessed to the point where they will be able to significantly affect the way we need and use animal models. Most likely, these techniques will be useful in identifying compounds likely to cause toxicity once placed in animals or humans and, thus, reduce the number of chemicals that need to be tested in animal models. Hopefully, they will also be able to tell which of us humans will probably experience adverse reactions to specific chemicals based on data likely to be obtained from both animal models and humans.

34.3.3.2  Electrophysiology Electrophysiological endpoints including electroencephalograms (EEGs) and event-related potentials such as the P300 wave (Pan et al., 1999) and the auditory brain stem response (Golub et al., 2004) are perfectly capable of detecting changes in nervous system activity as a result of exposure to occupational and other environmental insults in animal models (Burchfield et al., 1976; Pan et al., 1999). And, importantly, similar if not identical endpoints can also be obtained from a variety of species including humans. However, such changes are not always readily interpretable from the standpoint of functional outcomes and adverse consequences in humans. Additionally, especially with respect to the more targeted assessments like evoked electrical brain potentials,

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it can be argued that these analyses are restricted to specific pathways or brain areas and may miss effects in other areas that are not assessed using these approaches. In the case of pesticides, however, it may suffice simply to detect any effect since any deviation from the normal or baseline condition is unwanted.

34.3.3.3  Functional Measures and Imaging Major arguments that support the use of nonhuman primates as translational models for pesticide research do not stem from demonstrations of past successes using these animals in pesticide research since the number of such studies conducted in nonhuman primates is small and none have utilized the approaches currently available that make this model ideal. For example, in one study, groups of rhesus monkeys were exposed to different, relatively low doses of fenthion chronically for 1 year during which body weights, ophthalmologic exams, clinical chemistries, hematologies, urinalyses, and cholinesterase inhibition were periodically monitored: no treatment effects were reported (Coulston et al., 1978). (It remains unknown whether that same treatment regimen would have affected the function of the nervous system had sophisticated functional assessments be employed). Rather, supporting evidence for the use of the nonhuman primate model will be offered from recent observations that the analyses of complex brain function—as evidenced primarily by behavior—and the technical advances in imaging technologies have reached levels of sophistication that makes translational studies more possible in these animals than in others. The most comprehensive view of nervous system function can be obtained by observing behavior, primarily because it is the final common output of all of its components. When assessed thoroughly, a great deal of information can be had regarding the functional integrity of those structures supporting or participating in the behavior under observation. Behavioral observations are noninvasive and can be conducted repeatedly as long as the subject is able to respond. Obviously, in humans, observable behaviors include written and verbal behaviors, behaviors that are not obtainable from animal models. Nevertheless, comprehensive behavioral assessments of nonhuman primates, using the same or nearly identical instruments that can be used with humans, can be employed throughout their entire lifespan and these are the approaches of interest for the present discussion. Likewise, PET (positron emission tomography) and MRI (magnetic resonance imaging) images can be obtained repeatedly over extended periods of time in the same subjects. PET imaging approaches allow for the visualization of specific molecular or cellular targets. Specific tracers (chemicals labeled with positron emitters) can be used to target specific neurotransmitter or hormone receptors, specific cell types, and specific cellular processes such as apoptosis, neurotransmission, etc.

Chapter | 34  The Nonhuman Primate as a Translational Model for Pesticide Research

An excellent example of this approach can be found in studies of chronic manganese exposure in macaque monkeys (reviewed in Burton and Guilarte, 2009) where, using PET imaging approaches, it was demonstrated that manganese toxicity is associated with impaired release of dopamine in the striatum. These same animals also exhibited behaviors similar to those seen in people with Parkinson’s disease (Burton and Guilarte, 2009). MRIs can provide fine structural detail and both of these technologies can be utilized repeatedly providing the opportunity to gather important time-course information such as the speed of onset and duration of lesion development and recovery, the effectiveness of treatment, etc. Functional MRI approaches can, and have been, used to explore brain circuitry involved with executive function (described below) in humans (reviewed in Collette et al., 2006). MRI approaches can be used in any animal species including humans and translation is relatively easy given that the same kind of information is obtained using the exact same technology and instrumentation. Interpretation is made more difficult if the animal model does not have the same structures or systems as humans. Here again the nonhuman primate comes closest to sharing structures, systems, and temporal patterns of growth and development with humans. Some MRI studies to define the developmental trajectory of the structure of the infant human brain have already been completed (Knickmeyer et al., 2008) and similar studies should follow for nonhuman primates.

34.3.4  Nonhuman Primate Behavioral Concordance with Humans Concerning the observation of behaviors that are relevant or identical to those observable in humans, a variety of age-appropriate approaches and instruments have been developed for use in nonhuman primates—and other animals (see Sharbaugh et al., 2003)—for the purposes of modeling human-relevant behaviors in animals. These include procedures such as classical eye-blink conditioning in which aspects of primitive learning can be assessed; paired comparison tasks for assessment of visual recognition memory and habituation; the Fagan test of infant intelligence in which the tendency of infants to focus on novel stimuli (a measure that correlates with intelligence in humans) and others (see Sharbaugh et al., 2003). Use of identical or similar endpoints in risk assessment and hazard identification studies obviates the need for extensive interpretation and extrapolation, particularly if the endpoints are analogous, if not homologous, with those also obtainable in humans. Use of analogous and homologous behaviors also facilitates a comparison of cognitive function in humans and nonhuman primates (Roberts, 1996). As might be expected, the developmental stage at which subjects are assessed heavily dictates the kinds of approaches that can

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be used. The numbers and kinds of behavioral assessments that are appropriate for use in both humans and nonhuman primates far exceed those appropriate for use in other animals as exemplified in recent monkey studies (e.g., Golub, 2008; Grant and Rice, 2008; Isoherranen and Burbacher, 2008; Paule, 1995, 2001, 2008; Paule et al., 1988a, b; Schneider et al., 2008). Thus, the nonhuman primate provides opportunity for a much more comparable assessment of nervous system integrity than any other animal model.

34.3.4.1  Neonatal and Infant Assessments For the first few minutes or hours after birth, the health of newborn humans is often repeatedly assessed using the APGAR scale or score, which entails the observation of: Appearance (color); Pulse (rate); Grimace (face); Activity; and Respiration (rate and vigor). For nonhuman primates, these evaluations are basically identical with the use of the simian APGAR scale (Tarantal and Hendrickx, 1989). Later, during the neonatal period in humans, the typical instrument used in the assessment of health, temperament and other qualities is the Brazelton Newborn Assessment Scale or NBAS (Brazelton, 1973; Brazelton and Nugent, 1995). For nonhuman primates, a modified version of that tool has been developed (Schneider and Suomi, 1992). This 2-min assessment is useful for at least the 1st month of life in monkeys, can be administered repeatedly to the same subject, and covers a wide range of functional metrics including those that assess physical orientation, state control/emotionality, motor maturity, activity, orienting, neuromotor function, reflexes, and development. Since the visual, auditory, and tactile systems are involved in a variety of these measures, their status also contributes to the overall assessment score of this tool. Schneider and Suomi (1992) have used this approach to demonstrate the positive effects of environmental enrichment in rhesus monkeys. Others have also developed standardized neonatal assessments for nonhuman primates that have human analogues (Golub and Gershwin, 1984). More recently, a battery of behavioral assessments has been used to determine the outcomes of rhesus monkeys produced by assisted reproductive technologies (Sackett et al., 2006). This consists of neonatal reflexes, self-feeding ability, recognition memory, object concept attainment, simple discrimination learning, and reversal and learning set acquisition and can be conducted over the first month of life.

34.3.4.2  Beyond Infancy In older animals, approaches have incorporated electrophysiological measures such as the P300 wave, a positive deflection in an event-related electrical brain potential that is thought to represent the transfer of information to consciousness (Picton, 1992). Like other evoked potentials that can also serve as metrics of brain activity, the P300 wave

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is associated with the introduction of some kind of sensory stimulus (sound, light, touch, etc.) and the response of brain circuitry to that stimulus. Electroencephalograms (EEGs) are other metrics of brain function that can be monitored repeatedly in the same subjects. In fact, previous nonhuman primate studies on prolonged (18 months), low-level exposures to pesticides (DDT, dieldrin, parathion, and carbaryl) have shown significant effects on EEGs but the significance of those findings remains undetermined (Santolucito and Morrison, 1971). Other approaches such as classical behavioral procedures such as eye blink conditioning (Sharbaugh et al., 2003), social interactions (Riddick et al., 2009), and operant behaviors (Popke et al., 2001a,b) can also provide valuable insights into critical aspects of brain function for which interpretation is less problematic.

34.3.5  Automated Assessments 34.3.5.1  The National Center for Toxicological Research Operant Test Battery For over 20 years an automated battery of positively reinforced cognitive function tests, developed at the FDA’s National Center for Toxicological Research (NCTR), has been used extensively to assess the brain function of both nonhuman primates (e.g., Frederick et al., 1998; Paule et al., 1988b; Schulze et al., 1988, 1989) and humans, primarily children (e.g., Baldwin et al., 2004; Chelonis et al., 2000, 2002, 2004; Paule and Cranmer, 1990; Paule et al., 1988a) but also some young adults (Paule et al., 2004). This battery, initially described by Schulze et al. (1988), was designed specifically to assess aspects of motivation, visual discrimination, time perception, learning, and short-term memory, all functions thought to be critical for normal success in human life. Substantial effort has been put forth in the validation and interspecies application of this instrument, which is known as the NCTR Operant Test Battery (OTB). The term operant refers to the fact that the subjects must operate something in their environment (here, pressplates or response levers) in order to obtain a reinforcer (banana-flavored food pellet). The operations of specific manipulanda in relationship to the specific contingencies for reinforcement (task rules) are captured, stored, and summarized by computer. The OTB is a behavioral tool consisting of five timed tasks, or games, each designed to generate behavior that is thought to depend upon the specific brain functions mentioned above; the specifics of each task can be found in previous publications (e.g., Schulze et al., 1988, 1989). (a) Assessment of Motivation A Progressive Ratio (PR) task is used to assess motivation. Here, subjects must increase their lever-pressing behavior to continue to receive reinforcers: initially, the first treat

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or reinforcer only costs two lever presses, the second costs four, the third six, etc. In a relatively short period of time (10 min as typically used in the OTB), a metric of response strength or motivation can be obtained. In experienced subjects, daily performance of this task is very predictable for a given animal but can vary tremendously between animals. Thus, one monkey may generally press a maximum of 30 times for a food pellet, whereas a different animal may press upwards of 150 times. Response rates in this task correlate highly with number of reinforcers earned and serve as the key metric of motivation. Post-reinforcement pauses (time after reinforcer delivery) serve as metrics of manual dexterity and motor function since during these pauses, subjects are retrieving and eating the food pellet reinforcers obtained by lever pressing. (b) Assessment of Visual–Spatial Discrimination Visual and spatial discrimination is assessed using performance of a Conditioned Position Responding (CPR) task. Here three press-plates aligned horizontally serve as the manipulanda; all can be illuminated from behind with colored stimuli. Initially, the center press-plate is illuminated with either a red, yellow, blue, or green color. The subject acknowledges (observes) the stimulus by pressing it, after which it is immediately extinguished and the two side press-plates are illuminated white. If the center pressplate color had been red or yellow, a response (choice) to the left press-plate is reinforced. If the center press-plate had been blue or green, a right press-plate choice is reinforced. Thus, if red or yellow, choose left, if blue or green choose right. Observing and choice reaction times and choice accuracies are the primary data of interest. This task typically generates relatively high response rates and experienced animals easily complete the task (make 60 correct choices) in less than the 5 min allowed. Once performance of this task has been mastered (e.g., average percent correct choices 85%), the rules of reinforcement can be reversed such that right press-plate choices are correct for red and yellow stimuli and left press-plate choices are correct for blue and green stimuli. Reversal procedures can be employed in a variety of discrimination tasks, can be designed to assess any sensory modality (tactile, visual, auditory, olfactory, and gustatory) and, importantly, they can be used repeatedly, each time generating data concerning aspects of cognitive flexibility and/or perseveration (resistance to change). (c) Assessment of Time Perception Time perception in the OTB is measured using a Temporal Response Differentiation (TRD) task, which requires subjects to generate a response of a specific duration. In this case, the active response lever must be depressed for at least some minimal amount of time but no longer than some maximum amount of time. While the specific window

Chapter | 34  The Nonhuman Primate as a Translational Model for Pesticide Research

of targeted response durations can be defined by virtually any values, as used in the OTB the values are set at 10–14 s. Thus, correct lever hold durations will be at least 10 but no more than 14 s in duration. Typical response accuracies of greater than 60% are seen in experienced monkeys and, as used in the OTB, the maximum task time allotted is 20 min. A variety of time-related information can be gleaned from responding in this task. For example, the average lever hold duration of timed responses is thought to provide a metric of timing accuracy, whereas the response spread (deviation) around the mean is thought to provide a metric of timing precision. In addition, the peak height of the response duration plot (maximum number of reinforced durations made during a session) is thought to provide yet another measure of motivation. While there are numerous approaches to the assessment of time perception (see Paule et al., 1999b), the TRD task has been used extensively in nonhuman primates, particularly in the assessment of acute effects of psychotropic drugs (Paule 2001; Paule et al., 1999b). Shifts in mean lever hold durations are thought to reflect changes in the speed of an internal clocking mechanism: shifts to the left (shorter lever hold durations) indicating an increase in clock speed (8 s is perceived as 10 s) and shifts to the right indicating a decrease in clock speed. Decreases in response variation (standard deviations around the mean) are thought to indicate an increase in timing precision, whereas increases portend a decrease in timing precision. In clinical assessments of children with ADHD, stimulant medication has been shown to increase precision in this task (Baldwin et al., 2004). (d) Assessment of Learning Learning capability is assessed using the OTB’s Incremental Repeated Acquisition (IRA) task. In this task, subjects are required to learn a specific sequence of lever presses each test session. The sequence changes each session; thus, the animals cannot predict which sequence will be correct for any given session. Here, four response levers, all aligned horizontally, are used and there are six levels of task difficulty. Initially, subjects must learn which one of the four levers will produce a reinforcer. After mastery has been demonstrated for this one-lever sequence (e.g., 20 reinforcers have been earned), the task difficulty is increased such that two levers must be pressed in a specific sequence, with the last lever of the sequence being the same lever mastered at the easiest level of task difficulty (i.e., at the one-lever sequence level of difficulty). After mastery is demonstrated for the two-lever sequence, the task difficulty is incremented again to a three-lever sequence, and so on, up to a six-lever sequence or until the task times out (typically after 35 min). Key metrics of learning in this task include percent task completed, response accuracy (correct lever presses/total lever presses) and response rate for each level of task difficulty and overall (collapsed across all levels of difficulty) accuracy and response rate. Differential analyses

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of error types can also provide information as to learning strategies and typical learning curves (errors plotted versus number of correct sequences completed) can be generated at each level of task difficulty. (e) Assessment of Short-Term/Working Memory Short-term or working memory is assessed in the OTB using performance in a Delayed-Matching-To-Sample (DMTS) task. This type of memory can be assessed using DMTS tasks and a variety of other delayed response tasks that have been recently discussed (see Paule et al., 1998; Rodriguez and Paule, 2008). In the NCTR OTB, subjects are required to remember simple, white and black geometric symbols over relatively short periods of time. Here, the three press-plates used for the CPR task are also used. At the start of each trial, the center press-plate is illuminated with a sample stimulus (square, circle, triangle, etc.). The subjects observe this stimulus and respond to it (press the plate), it is immediately extinguished and a randomly chosen recall delay (e.g., from 1 to 32 s) begins. At the end of the recall delay, all three press-plates are illuminated, each with a different geometric symbol, one of which matches the sample stimulus for that trial. A response (choice) to the stimulus that matches the sample stimulus results in the delivery of a banana-flavored food pellet. Plots of choice accuracy by recall delay provide metrics of working memory: the slope of the line indicates rate of forgetting and the Y intercept provides information about both encoding (formation of a memory trace) and attentional processes. Increases in the negative slope of the decay curve suggest a more rapid memory decay, whereas decreases indicate longer retention. Decreases in the Y intercept suggest a decrement in either encoding processes, attention, or both. By examination of choice latencies as a function of recall delay, aspects of attention can be further isolated: as recall delays increase, the likelihood of distraction increases; thus, additional metrics of the attentional effects of experimental manipulations can be obtained. The duration of the DMTS task as used in the OTB is typically 30 min. (f) Minimization of Assessment Bias The apparatus used to administer the NCTR OTB is completely automated, obviating the need for tester–testee interaction and, therefore, eliminating the bias that can sometimes influence similar assessments that rely on significant interaction with the test giver and the test taker [e.g., the Wisconsin General Test Apparatus or WGTA, originally described by Harlow, (1959)]. Automation also minimizes technical support needs and allows for the assessment of relatively large numbers of subjects in a relatively short period of time: in our laboratory, 90 animals have been routinely assessed daily for a period of many months. Thus, relatively high throughput assessments are feasible.

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(g) Model Validation Concordance with Human Findings  Importantly, and unlike many other animal models and test systems, OTB performance in the pubescent and adult nonhuman primate has been validated as a surrogate for human psychotropic drug effects and details of the process can be found in earlier publications (Paule, 1998, 2001, 2005, 2007). Acute and chronic drug effects on NCTR OTB behavior in the monkey model are highly concordant with/predictive of drug effects in humans. For example, when adult rhesus monkeys are given the active ingredient in marijuana (delta-9-tetrahydrocannabinol; THC), performance of the OTB timing task indicates that these animals overestimate the passage of time, releasing the response lever too early (e.g., 8 s is perceived as 10 s), an effect that replicates human responses to THC. When marijuana smoke itself is administered to rhesus monkeys, the OTB findings indicate that short-term memory is very sensitive to disruption, again a finding parallel to that seen in humans. When pubescent rhesus monkeys were used as a model of human teenage marijuana smokers, chronic exposure to relevant doses resulted in an amotivational syndrome characterized by much less vigorous responding in the OTB motivation task in marijuanausing animals than in control animals. An amotivational syndrome has also been reported in several studies of teenage and young adult human marijuana users (discussed in Paule, 2001). Additional examples of the ability of nonhuman primate OTB behavior to predict acute drug effects in humans can be found in Paule (2001). Profiling Psychotropic Drug Effects  In addition to making direct human-to-monkey comparisons of drug or chemical effects, the OTB can also be used to characterize the sensitivity of different brain functions to chemical agents. For example, dose–response information can be gathered from all five OTB tasks (PR, CPR, TRD, IRA, and DMTS) and their relative sensitivities can be determined. Thus, in animals exposed to THC, the timing (TRD) task is most sensitive to its effects followed by the short-term memory (DMTS), learning (IRA), and color and position discrimination (CPR) tasks, followed by the motivation (PR) task, yielding the profile: TRD  IRA  DMTS  CPR  PR. For every compound assessed so far, a different behavioral profile has been identified, even for drugs within the same pharmacological class such as the stimulants (methylphenidate, amphetamine, and cocaine) (Paule, 2001). This is noteworthy since it indicates that the OTB can detect subtle differences between chemicals thought to act via very similar mechanisms. While no pesticides have yet to be tested using this paradigm, data on the acute effects of physostigmine, a compound that shares acetylcholinesterase inhibition properties with the organophosphates, indicate the following OTB profile: TRD  DMTS  PR  IRA (Frederick et al., 1995). The observation that each task was affected at the same dose(s) of physostigmine suggests that a mechanism

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common to all tasks underlies these findings. Given that all tasks have similar motor/response requirements, it is likely that the motoric effects associated with cholinesterase inhibition are responsible for these findings. Importantly, others have indicated that cholinesterase inhibition in humans is not likely to account for the cognitive/neuropsychiatric effects seen with chronic exposure to OP pesticides (Pancetti et al., 2007; Salvi et al., 2003). In addition, others studying the cognitive effects of weeks of exposure to diisopropylfluorophosphate in rhesus monkeys showed that even in the presence of significant erythrocyte cholinesterase inhibition (75%) there was no effect on DMTS performance (Prendergast et al., 1998). Indeed, the chronic effects of agents are often different from their acute effects, as noted in our own studies where the short-term memory task was most sensitive to the acute effects of marijuana smoke (Schulze et al., 1989), yet motivation behavior—the behavior least affected by acute exposure—was the most sensitive to chronic exposure (Paule et al., 1992). Direct Comparison of Monkey Versus Human Per­formance  Studies have also been conducted to compare, directly, nonhuman primate OTB behavior with that of humans (children) (Paule et al., 1990). In general, the OTB performance of experienced young adult monkeys is indistinguishable from that of children from ages 4 to 13 years, depending upon OTB task and endpoint. For example, DMTS accuracies and memory decay functions in 4-year-old children are nearly identical to those for young adult rhesus monkeys (M.G. Paule, unpublished observations). On the other hand, DMTS choice response latencies for young adult monkeys are no different than those for 13-year-old children (M.G. Paule, unpublished observations). Thus, while working memory, per se, of monkeys is on par with that of younger children, the speed with which monkeys can respond in this task is equivalent to that of much older children. These observations demonstrate clear interspecies concordance of OTB behavioral output. Demonstration of Clinical Relevance in Humans: the IQ Equivalent  With respect to human relevance, it has been shown that the endpoints of several OTB tasks when performed by children are highly and significantly correlated with the gold standard clinical measure of intelligence: IQ score (Paule et al., 1999a). Subsequently, computer algorithms were constructed such that human IQ scores can be predicted relatively well from OTB data alone and, by analogy, metrics of monkey intelligence can also be inferred (Hashemi et al., 1994).

34.3.5.2  Additional Automated Approaches (a) The Cambridge Neuropsyhcological Automated Test Battery and Other Touch Screen Approaches While the NCTR OTB has been used extensively in nonhuman primate studies, other similar systems and

Chapter | 34  The Nonhuman Primate as a Translational Model for Pesticide Research

approaches have also been developed. A touch screen apparatus, initially developed for human use, the Cambridge Neuro­psychological Test Automated Battery (CANTAB; Fray et al., 1988; Luciana and Nelson, 2002; Purcell et al., 1997) has also been used in nonhuman primate studies. This battery, for which human norms have been specified for some age groups (e.g., Robbins et al., 1994, 1998) also consists of several operant tasks designed to assess specific aspects of cognitive function (see Chudasama and Robbins, 2006 for list of comparable tests of cognition in rats, monkeys and humans). Many of the tests in the CANTAB are nonverbal, like those in the OTB and, thus, are also appropriate for use in animals, in nonverbal human subjects, and subjects from different cultures. The CANTAB contains tests designed to monitor aspects of general memory and learning, working memory, executive function (planning, cognitive flexibility, selective attention, response inhibition, etc.), visual memory, attention and reaction time, decision making, and response control. Thus, the CANTAB represents an instrument designed initially for human use that, because it is automated, is operant in nature, and uses nonverbal stimuli is finding utility in nonhuman primate testing. The NCTR OTB on the other hand was designed for use in the nonhuman primate laboratory but is finding use in the clinical setting because of its relevance, ease of administration, and appropriateness for human subjects. Use in Animal Studies  Performance norms for rhesus monkey acquisition and performance of some CANTAB tasks have been obtained (Weed et al., 1999) and some acute drug studies have been completed in monkeys (Taffe et al., 1999). Studies using the CANTAB have also assessed the effects of chronic, peripubertal exposure of female rhesus monkeys to the estrogenic pesticide, methoxychlor, on aspects of cognition (Golub 2002; Golub et al., 2004). Performance of a Delayed-Non-Matching-To-Sample task indicated that methoxychlor delayed task acquisition and caused working memory difficulties, albeit at relatively high doses; effects were noted at 50 but not 25 mg/kg/day. There was also an effect of methoxychlor to alter growth patterns and ovarian cyclicity (Golub et al., 2003). Spinelli et al. (2006) have used marmoset performance of some CANTAB tasks (five-choice serial reaction time and concurrent delayed matching to position, for the assessment of attention and working memory, respectively) to study the acute effects of nicotine and scopolamine and reported slightly enhanced performance after nicotine and disrupted performance after scopolamine. Still others have used marmosets to study the acute effects of sarin on both EEG and CANTAB discrimination behavior and reported no or minimal effects ((Pearce et al., 1999). In other studies on rhesus monkeys, performance of Delayed Matching-to-Sample and Self-Ordered Spatial Search tasks was disrupted by the acute administration of ketamine, a noncompetitive inhibitor of the N-methyl-D-aspartate (NMDA) glutamate receptor,

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whereas Progressive Ratio responding—motivation—was not (Taffe et al., 2002). (b) Other Touch Screen Systems Recently, others have begun to adopt automated approaches for use in nonhuman primates with a view towards obtaining data from quite young animals (Mandell and Sackett, 2008a,b). Typically, the OTB has been used in monkeys only after weaning at 6–7 months of age but others have recently employed touch screen technology in attempts to begin automated assessments in younger animals (3–4 months of age) and to compare computerized assessments with those obtained using older methods (Mandell and Sackett, 2009). Some success was realized in younger animals, suggesting that operant type assessments will be applicable at even younger ages as methods development progress.

34.4  Overview and summary The upshot of this discourse is to demonstrate that our ability to effect comprehensive, human-relevant assessments of the state of the nonhuman primate brain in even very young monkeys has evolved nicely and thoroughly over the past two to three decades and it has done so with a view toward clinical relevance and translational capabilities. Many of the current instruments that are used in the nonhuman primate laboratory are applicable, if not identical, to those used clinically. Thus, issues of cross-species extrapolation can be minimized or eliminated. The demonstration that the monkey model can, at least in cases where comparable data exist, predict the acute and chronic effects of psychotropic drugs on rather sophisticated aspects of brain function in humans confirms its utility. In studies that exemplify the approaches to study design championed in this chapter [i.e., chronic exposures in young monkeys to doses of chemicals that are relevant to humans while repeatedly monitoring rel­ evant aspects of brain function (e.g., OTB behavior)], significant adverse effects on specific brain functions have been noted to occur in the absence of effects on those endpoints typically monitored in toxicology studies (Paule et al., 1992; Popke et al., 2001a,b, 2002; Slikker et al., 1991). The insensitive endpoints used in these studies include the usual suspects: body weights, clinical chemistries, hematologies, ophthalmological exams, and urinalyses. Thus, particularly in studies where the agent of interest is known to at least enter brain tissue, it has become clear that the typical approaches to the assessment of toxicity can be inadequate, if not misdirected. That is not to say that the usual endpoints are not informative: they have clearly demonstrated that otherwise healthy, normal-appearing animals can manifest significant deficits in important aspects of brain/cognitive function. There are relatively few circumstances in which non­ human primates have been used in pesticide research,

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particularly with respect to determining functional consequences under real-world exposure scenarios. For example, in one study on DDT carcinogenicity, cynomolgus and rhesus monkeys were exposed to high doses via their diet for over 10 years, at the end of which 25% showed severe tremors and histological evidence of CNS and ­spinal cord abnormalities; fatty changes in liver were also noted, but detailed behavioral analyses were not carried out (Takayama et al., 1999). There are, however, ample examples of the utility of the model in other areas of research relevant to this topic. In fact, a recent book has been devoted to nonhuman primate models of children’s health and developmental disabilities (Burbacher et al., 2008). Clear examples of how nonhuman primates have served as valuable resources in this regard can be found here with reviews of the nonhuman primate lead, methylmercury, and PCB literature demonstrating the contributions made so far (Grant and Rice, 2008). It is also important to note that social behaviors, which can be extremely complex in nonhuman primates and clearly have human counterparts, can also serve as surrogates for human behaviors and, thus, endpoints of use in studying the effects of pesticides. Since the procedures needed to conduct these kinds of studies are not easily automated, are relatively labor-intensive, and require environments not easily configurable in typical animal laboratory settings, they have not been highlighted here. Nevertheless, these approaches are of considerable value and certainly have face validity when it comes to extrapolation of findings to humans. While the costs associated with nonhuman primate studies is considerable, it can be argued that untold millions of dollars have already been spent supporting studies in probably inferior animal models, using irrelevant exposure paradigms given at the wrong life stages. For new or untested compounds, our nonanimal and nonprimate animal models should be used to identify those compounds most likely to give a toxic signal related to nervous system function. For these and other high-priority chemicals, such as those of very high volume production, then it will be very reasonable to use the nonhuman primate to provide insight into the effects likely to be encountered in humans.

Acknowledgments The opinions and interpretations presented herein are solely those of the author and are not intended to represent those of the U.S. Food and Drug Administration.

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Section VI

Pesticides Disposition

(c) 2011 Elsevier Inc. All Rights Reserved.

Chapter 35

Introduction to Pesticide Disposition Ernest Hodgson North Carolina State University, Raleigh, North Carolina

It should be emphasized that, although pesticides and their use have many positive attributes, in terms of their interac­ tions with living organisms, pesticides are xenobiotics and are processed in the same way as other xenobiotics such as clinical drugs and industrial chemicals. It should also be emphasized that toxicity is not due to a single defining molecular event or interaction but rather a cascade of events beginning with exposure and culminating with the expression of one or more toxic endpoints. This cascade (Figure 35.1) includes adsorption, distribution, metabolism (both detoxication and activation), distribution of metabolites, interaction with cellular macromolecules (such as RNA, DNA, and proteins), repair, and excretion. The processes involved may be reversible to a greater or lesser extent, they may be alternate pathways, and they may be modified by chemical and physiological interac­ tions. Thus, exposure to a toxicant does not inevitably lead to a toxic endpoint, metabolism, excretion or repair may render the original exposure without effect (Hodgson, 2004). Finally, these processes and the genes, enzymes,

transporters, receptors, etc. involved are all subject to con­ siderable variation with cell type, organ, individual, spe­ cies, and strain. The aspects covered in this section include adsorption, distribution, metabolism, and excretion and are collectively known as disposition. In this presentation, they are divided into six chapters written by the following four authors. Dr. Ronald Baynes, Associate Professor of Pharma­ cology in the College of Veterinary Medicine at North Carolina State University. Dr. Baynes received his BS (with Honors) from the University of the West Indies (Cave Hill Campus), his DVM (with Honors) from Tuskegee University, his MS from the University of Georgia, and his PhD from North Carolina State University. His research is focused on using QSAR modeling approaches to under­ standing the physicochemical factors influencing dermal absorption of pesticides and formulation additives that cause occupational irritant dermatitis. Dr. Kelly J. Dix, the author of Chapter 24 in the 2nd edition of the Handbook of Pesticide Toxicology, is no

Figure 35.1  Chemical toxicity: a cascade of events. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

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longer involved in toxicological pursuits or with the Handbook. However, her previous chapter was the progen­ itor of the current Chapter 39 and, as a result, contributed to it. Her contribution is much appreciated. Dr. Ernest Hodgson is a Distinguished Professor Emeritus in the Department of Environmental and Mole­ cular Toxicology at North Carolina State University. Dr. Hodgson was awarded his BSc with Honors by the University of Durham (UK) and his PhD by Oregon State University. He has been interested in the metabolism of pesticides for many years and, more recently, has focused on the human metabolism of pesticides. Dr. Hodgson is also known for the publication of two widely accepted textbooks of toxicology. Dr. Jim E. Riviere is the Burroughs Wellcome Fund Distinguished Professor of Pharmacology and Director of the Center for Chemical Toxicology Research and Pharmacokinetics, College of Veterinary Medicine, North Carolina State University in Raleigh, NC. Dr. Riviere

received his BS (summa cum laude) and MS degrees from Boston College, his DVM and PhD in pharmacology as well as a DSc (hon) from Purdue University. He is an elected member of the Institute of Medicine of the National Academies, serves on its Food and Nutrition Board, and is a fellow of the Academy of Toxicological Sciences. His current research interests relate to the development of ani­ mal models; applying biomathematics to problems in toxi­ cology, including the risk assessment of chemical mixtures, pharmacokinetics, nanomaterials, absorption of drugs and chemicals across skin; and the food safety and pharmaco­ kinetics of tissue residues in food-producing animals.

Reference Hodgson, E. (2004). Introduction to toxicology. In “A Textbook of Modern Toxicology” (E. Hodgson, ed.), 3rd ed. John Wiley and Sons, Hoboken, New Jersey.

Chapter 36

Introduction to Biotransformation (Metabolism) Ernest Hodgson North Carolina State University, Raleigh, North Carolina

36.1  Introduction Williams (1959) first suggested that the metabolism of xenobiotics generally occurs in two phases. The word xenobiotic, however, was coined later, in the mid 1960s, by Dr. Howard Mason (1964) to serve as a collective noun including any chemical to which an organism is exposed and is extrinsic to the normal metabolism of that organism. Thus it includes pesticides, occupational chemicals, clinical drugs, drugs of abuse, deployment-related chemicals, etc., and is a particularly useful term when discussing metabolic pathways and enzymes that have substrates in several of these use classes. Phase I involves predominantly oxidations, reductions, and hydrolysis and serves to introduce a polar group into the molecule. Phase II, consisting primarily of conjugation reactions, involves the combination of the products of phase I reactions with one of several endogenous molecules to form water-soluble, and hence excretable, products. A number of books review the biotransformation of xenobiotics, either in general or of particular chemical or use classes (e.g., Hodgson and Levi, 1994, 1997; Jakoby, 1980; Jakoby et al., 1982; Klaassen, 2001; Smart and Hodgson, 2008; Wilkinson, 1976; Williams, 1959). Many treatments of pesticides (e.g., Chambers and Carr, 1995; Ecobichon, 2001; Hodgson and Meyer, 1997; Hodgson et al., 1995; Kulkarni and Hodgson, 1984a,b; Rose et al., 1999) include considerations, not only of pesticide metabolism, but also of the significance of metabolism in the toxicity of pesticides to target and non-target species. In the past, most emphasis has been placed on microsomal cytochrome P450 (CYP)-dependent oxidations and reductions of pesticides. These are discussed below and in Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

subsequent chapters. However, more recently, much has been learned of the roles of other phase I enzymes such as flavin-dependent monooxygenases (FMO), hydrolases, and epoxide hydrolases, and of cooxidation during prostaglandin synthesis. Considerable emphasis has also been placed on the phase II conjugation reactions as they apply to pesticide metabolism. By definition, microsomal enzymes are those found in the high-speed particulate microsomal, fraction of tissues following homogenization and differential centrifugation. The terms microsomal fraction and microsomes refer to a biochemical preparation and do not correspond to any particular cell structure. However, the major component consists of membranous vesicles derived from the endoplasmic reticulum and its constituent ribosomes. The microsomal fraction consists primarily of rough (with ribosomes) and smooth (without ribosomes) vesicles that correspond to rough and smooth endoplasmic reticulum. The specific activity of smooth microsomes is generally higher than that of rough microsomes for the metabolism of xenobiotics. However, even though rough and smooth microsomes can be separated by density gradient centrifugation, this is generally not done in pesticide metabolism investigations. The majority of studies focusing on pesticide metabol­ ism and the regulation of pesticide-metabolizing enzymes have been conducted in experimental animals, primarily rodents. However, there has been an increase in information about human enzymes, especially the CYP isoforms. Much of this information has been gained through the use of specific substrates, antibodies, human hepatocytes, human cell fractions, and recombinant human enzymes. Studies with human CYPs have become more common and have 865

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demonstrated that xenobiotic metabolism and the regulation and expression of xenobiotic-metabolizing enzymes may be quite different in humans and in experimental animals. Such differences make the extrapolation of metabolism studies from experimental animals to humans difficult. It is only as we learn to understand these differences that we can make more accurate and realistic extrapolations to humans.

36.2  Xenobiotic-metabolizing enzymes 36.2.1  Cytochrome P450 Monooxygenases Although several enzymes acting in concert may be required for xenobiotic degradation or activation, the initial reaction usually involves a microsomal phase I enzyme catalyzing an oxidation reaction. Reduction reactions, although they may also occur, are relatively uncommon. These enzymes include many of the isoforms of CYP active in the CYP-dependent monooxygenase system, as well as FMO isoforms. The overall aspects of the biochemistry and molecular biology of the CYP system are discussed in some detail by Zeldin and Seubert (2008). Many different pesticide monooxygenation reactions are attributed to CYPs, including epoxidation (e.g., aldrin), Ndealkylation (e.g., alachlor, atrazine), O-dealkylation (e.g., chlorfenvinphos), S-oxidation (e.g., phorate), and oxidative desulfuration (e.g., parathion) (Hodgson, 1982–1983; Kulkarni and Hodgson, 1980, 1984a,b). They are discussed in detail in Chapter 38.

Currently the CYP superfamily comprises, in all taxa, over 7000 genes classified into 781 gene families. Of the 110 animal CYP families, 18 are found in vertebrates (Nelson et al., 2008). The total number of functional CYP genes in any single mammalian species is thought to range from 60 to 200 (Gonzalez, 1990). In vertebrates, most CYP families encode proteins involved primarily in specific endogenous functions (i.e., steroid hormone biosynthesis and metabolism). Other families encode proteins that appear to have more to do with the oxidation of exo­ genous compounds, such as pesticides, and often display a broad range of substrate specificity (Bogaards et al., 1995; Nebert et al., 1989). The numerous polymorphisms found in these CYP genes are of considerable importance in the metabolism of xenobiotics, including pesticides. A polymorphism is defined as an inherited monogenetic trait that exists in at least two genotypes (two or more stable alleles) and is stably inherited. They arise as mutational events, if the mutation is in the coding region a variant protein is expressed, thereby affecting the rate and/or extent of metabolism and, potentially, the ratio of different metabolites and the extent of activation vis-à-vis detoxication. The details of the interaction of CYPs with xenobiotics have been the subject of intense study for some time, although in these studies clinical drugs have been utilized to a much greater extent than pesticides. Despite the wide range of amino acid sequence homology among CYP isoforms, all these isoforms have a remarkably similar reaction mechanism (Figure 36.1). CYP monooxygenation reactions all involve the reduction of one atom of molecular oxygen

ROH RH

Fe3+ 9

Fe3+ ROH

Fe3+ RH

1 8

1 e-

FeOH3+ R•

2

10 R’O R’OOH RH

7 FeO3+ RH

NADPH-P450 reductase

Fe2+ RH

H 2O

3

O2

6 Fe2+ -O2 RH FeII - OOH

RH

4 5 1 eFe2+

H+

-O2 -RH

NADPH-P450 reductase or NADHCytochrome b5 reductase

Figure 36.1  Catalytic cycle for cytochrome P450 (CYP)-catalyzed monooxygenation reactions. (From Zeldin and Seubert, 2008; Adapted from Guengerich FP, Chem. Res. Toxicol. 2001, 14: 611–650.)

Chapter | 36  Introduction to Biotransformation (Metabolism)

to water and the incorporation of the other oxygen atom into the substrate. The electrons involved in the reduction of CYP are transferred from NADPH by the NADPH-cytochrome P450 oxidoreductase while, in some cases, the second electron may be derived from NADH via cytochrome b5. Reviews of pesticide studies include Hodgson and Kulkarni (1974), Hodgson (1974), and Kulkarni and Hodgson (1980, 1984a,b). Studies of spectral interactions of pesticides with CYP, interactions that may be indicative of the ability to act as substrate or inhibitor, have also been carried out (Mailman and Hodgson, 1972; Mailman et al., 1974). Studies using specific isoforms (Hodgson et al., 1998; Levi and Hodgson, 1984, 1988) indicate that even in the same organ of the same species particular pesticides are metabolized at different rates by different CYP isoforms. The specificity of different isoforms for pesticide substrates is an area of current interest. Due to the availability of recombinant human isoforms, these studies can now be carried out on human enzymes as well as those from experimental animals. In an early study of fenitrothion metabolism by mouse liver utilizing four constitutive and two induced CYP isoforms, Levi et al. (1988) showed all isoforms produced both the cresol detoxication product and the oxon. However, there were significant differences both in overall activity and in the oxon/cresol ratio. The most active isoform, induced by phenobarbital and now known as CYP2B10, was also active in the metabolism of parathion and methyl parathion and in all cases produced significantly more oxon and detoxication products. Human CYP3A4 was shown to be most active in the metabolism of parathion, although CYP1A2 and 2B6 also showed activity (Butler and Murray, 1997). In studies on the metabolism of chlorpyrifos by human CYPs (Tang et al., 2001), it has been shown that CYP2B6, CYP2C19 and CYP3A4 are all active, CYP2B6 producing an excess of chlorpyrifos oxon and CYP2C19 an excess of detoxication products, while CYP3A4 produces both in approximately equal quantities. Studies of triazine herbicides in mice (Adams et al., 1990) and rats (Hanioka et al., 1999) as well as in rats and pigs (Lang et al., 1996) suggested a broad lack of isoform specificity for these substrates. However, Lang et al. (1997) showed that, in humans, CYP1A2 appeared to be the principal, if not the only, isoform responsible for triazine herbicide oxidation. Inui et al. (2000) expressed human CYP1A1, CYP2B6, and CYP2C19 in potatoes and produced resistance to several herbicides, including atrazine, in the host plants, presumably by enabling the plants to metabolize the herbicides. In studies of chloroacetanilide herbicides (Coleman et al., 1999), it was shown that human CYP3A4 was responsible for the initial O-dealkylation of alachlor. Subsequent studies (Coleman et al., 2000) extended these studies to acetochlor, butachlor, and metachlor. In all cases, CYP3A4 was the most active human isoform, although CYP2B6 also had some activity.

867

One of the significant features of many of the microsomal CYPs is their inducibility by xenobiotics; thus, stimulation of the metabolism of a chemical by prior administration of the same or another chemical is often taken as presumptive evidence of its metabolism by microsomal enzymes. For example, in mice pretreated with phenobarbital there is an increase in phorate metabolism, suggesting that CYP isoforms, such as CYP2B or CYP3A forms, may be important in the metabolism of similar pesticide substrates (Kinsler et al., 1990). Metabolic interactions involving enzyme induction and/or inhibition are discussed in Chapter 40. CYP-dependent reactions, as they involve pesticides, are summarized in Figure 36.2 and considered in detail in Chapter 38. A more mechanism-based classification of CYP-catalyzed xenobiotic oxidations is that of Guengerich and MacDonald (1984). They classified such reactions into six general categories: 1. Carbon hydroxylation: the formation of an alcohol at a methyl, methylene, or methine position 2. Heteroatom release: the oxidative cleavage of the heteroatom part of a molecule resulting from a hydroxylation adjacent to the heteroatom that generates a geminal hydroxy heteroatom-substituted intermediate such as a carbinolamine, halohydrin, hemiacetal, hemiketal, or hemithioketal. (This intermediate then collapses to release the heteroatom and form a carbonyl compound.) 3. Heteroatom oxygenation: the conversion of a heteroatomcontaining substrate to its corresponding heteroatom oxide as in the formation of N-oxides, sulfoxides, or phosphine oxides 4. Epoxidation: the formation of oxirane derivatives of olefins or aromatic compounds 5. Oxidative group transfer: a type of reaction that involves a 1,2-carbon shift of a group with the concurrent incorporation of oxygen to form a carbonyl at the C1 position 6. Olefinic suicide destruction: inactivation of the heme of P450 by an enzyme product

36.2.2  Flavin-containing Monooxygenase The microsomal flavin-containing monooxygenase was known for a number of years as an amine oxidase but was subsequently shown to be also a sulfur oxidase and a phosphorous oxidase. Like CYP, the FMO is a microsomal enzyme, a monooxygenase requiring NADPH and oxygen, and exists as multiple isoforms in various tissues. However, FMO, unlike CYP, catalyzes only oxygenation reactions, has more specific substrate requirements, and is not known to be subject to induction or inhibition by xenobiotics, apart from competitive inhibition by alternate substrates (Kulkarni and Hodgson, 1984a,b; Ziegler, 1980). The mechanism of catalysis is also distinct (Figure 36.3) in that electrons are transferred directly from NADPH, and not via an NADPH-reductase. Also, because the

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868

Oxidation N-Dealkylation

R N

C2H5 C2H5

Phosphamidon

O-Dealkylation

O C2H5O P OX C2H5O Chlorfenvinphos

Epoxidation Stable epoxide formation

* *

R N

O HO P OX C2H5O

*

Heptachlor

CH3CHO Acetaldehyde

CH3CHO

�

Acetaldehyde

Desmethylchlorfenvinphos

*

*

�

C2H5

Desethylphosphamidon

* * *

H

*

*

H O

* * *

* = CI

H *

Heptachlor epoxide H O

Arene oxide formation Naphthalene OCONHCH3

H

Naphthalene 1,2-epoxide OCONHCH3

Ring hydroxylation (usually via arene oxide formation)

4- Hydroxyl-1-naphthyl n-methyl carbamate Carbaryl OCONHCH3

OH OCONHCH2OH

Side chain hydroxylation

1-Naphthyl N-hydroxymethyl carbamate Carbaryl S

Sulfoxidation

(C2H5O)2PSCH2SC2H5

Phorate Figure 36.2a  Chemical reactions in pesticide metabolism.

formation of the hydroperoxyflavin form of the enzyme precedes interaction with the substrate, maximum velocity (Vmax) for a particular FMO isoform is constant for all substrates, although the Michaelis constant (Km) can vary from one substrate to another. CYP isoforms, on the other hand, show variations from one substrate to another in both Vmax and Km. The FMO is found in highest levels in the liver, but is also found in significant levels in the lung and kidney. Recent studies have identified five forms of FMO (FMO1–FMO5), which are differentially expressed with respect to species and tissue (Lawton and Philpot, 1995; Lawton et al., 1994). Each species that has been examined by analysis of genomic DNA appears to contain the same set of FMO genes (Lawton et al., 1994). Although the FMO family possesses multiple isoforms, the number

S

O

(C2H5O)2PSCH2SC2H5 Phorate sulfoxide

of such forms is small compared to that of the CYP superfamily. While the FMO isoforms are catalytically similar, marked differences do exist in substrate specificity. The importance of the FMO in pesticide metabol­ism was established when it was discovered that the FMO oxidizes a variety of thioether-containing pesticides (Cherrington et al., 1998a,b; Hajjar and Hodgson, 1980, 1982a,b; Levi and Hodgson, 1992; Smyser et al., 1985; Tynes and Hodgson, 1985). It has since been shown that the FMO is capable of oxidative desulfuration (oxon formation) of certain phosphonate insecticides such as fonofos through a mechanism distinct from that of oxon formation by CYPs (Smyser and Hodgson, 1985; Smyser et al., 1985) as well as the oxidation of pesticides from a number of different chemical classes (Tynes and Hodgson, 1985).

Chapter | 36  Introduction to Biotransformation (Metabolism)

NHCOCH(CH3)2

NHCOCH(CH3)2 OH Hydroxy IPC

IPC

N-Oxide Formation

N

N N

N

CH3

R

O. CH2 O

R

R

Desulfuration/or dearylation

R

O. CHOH O

R

S (C2H5O)2P O

O

CH3

Nicotine-1’-N-Oxide

Nicotine Methylenedioxy ring cleavage

Figure 36.2B�  (Continued)

O

O N-Oxidation

869

R

OH

R

OH�HCHO Catechol

O. C: O

R

Complexes with Fe+2 of cytochrome P450 O (C2H5O)2POH Diethyl phosphate

NO2

Parathion

HO

NO2 �

p-Nitrophenol

O S P

S

(C2H5O)2POH Diethylphosphorothioate

O (C2H5O)2P O

NO2 � [S]

Paraoxon Reduction Reduction of nitro group

S (C2H5O)2P O

S NO2

Cl

CH CCl3

NH2

Aminoparathion

Parathion

Dechlorination

(C2H5O)2P O

Cl

DDT

Reviews include Hodgson (1982–1983), Kulkarni and Hodgson (1984a,b), and Hodgson et al. (1998). FMO isoform specificity in pesticide metabolism is also being investigated. For example, Cherrington et al. (1998a,b) showed that, in the mouse, FMO1 metabolizes phorate to phorate sulfoxide but FMO5 is without activity. A recent review (Hodgson et al., 2008) covers aspects of xenobiotic metabolism by the FMO.

36.2.3  Other Phase I Reactions

Cl

CH CHCl2

Cl

DDD

are a family of enzymes known to exist both in the endoplasmic reticulum and in the cytosol. The overall attributes of these enzymes are discussed in Arand et al. (2005) and Hodgson et al. (2008). Epoxide hydrolases are known to attack xenobiotics of many classes, including some pesticide substrates, although these reactions are subsequent to the initial formation of epoxides. Examples include naphthalene 1,2-oxide and the 3,4-and 5,6-epoxides of carbaryl (Dorough and Casida, 1964) and tridiphane (Magdalou and Hammock, 1987).

36.2.3.1  Epoxide Hydrolases

36.2.3.2  Prostaglandin Synthetase

Epoxide rings of certain alkene and arene compounds are hydrated enzymatically by epoxide hydrolases to form the corresponding trans-dihydrodiols. The epoxide hydrolases

Prostaglandins are synthesized in mammals via a reaction sequence starting with arachidonic acid as substrate. During the second, or peroxidase, step of prostaglandin

Hayes’ Handbook of Pesticide Toxicology

870

Reduction of a double bond

R

C C

R

H CI

R C C CI

2H

H H DDMS

DDMU

Hydration of a double bond

R R

C C

H H

Figure 36.2c  (Continued)

R H +

R H +

R C C OH

HOH

H H DDOH

DDNU Hydrolysis O Phosphate ester hydrolysis

O

R2PO(S)R’

R2POH

Most organic phophorus esters

Acid

HO(S)R’ Alcohol

O

O Amide cleavage

+

RCNHCH3 Dimethoate (showing part of side chain)

RCOH

NH2CH3

+

O-O-Dimethyl-S-carboxylmethyl phosphorodithioate

Methylamine

O

O Thioester cleavage

Deamination

CH3(CH2)2SCN(C2H5)2

CH3(CH2)SH

Pebulate

Propyl mercaptide

O

O

CH3O P NHCH3 CI

CH3O P OH CI

+

+

HOCN(C2H5)2 Acid

NH2CH3 Methylamine

C(CH3)2 Rulene®

C(CH3)2 Deaminomethyl Rulene®

Desaturation 1,2,3,4,5,-Hexacyclorocyclohexane

Hexachlorocyclohexene

synthetase action, xenobiotics can be cooxidized to yield products similar to those formed by various isoforms of CYP (Eling et al., 1983; Hodgson et al., 2008; � Marnett and Eling, 1983). A number of pesticides (e.g., aminocarb, parathion) have been shown to act as substrates. These reactions may be important in extrahepatic tissues low in CYP and high in prostaglandin synthetase, such as the seminal vesicle and the inner portion of the medulla of the kidney.

including pesticides. For example, dimethoate is detoxified by amidase activity and the selective toxicity of malathion is due, in large part, to the presence in mammals of carboxyl­ esterases not widely distributed in insects. These enzymes are known from both microsomes and the soluble cytoplasm but are more commonly found in the latter. It appears likely that, in most cases, amidase and esterase activities are diff­ erent activities of the same enzymes with one or the other activity predominating (Satoh, 1987).

36.2.3.3  Hydrolases and Amidases

36.2.3.4  DDT-Dehydrochlorinase

Hydrolase and amidase activities (Hodgson, 2008) are known to be important in phase I metabolism of xenobiotics,

In the early 1950s, it was demonstrated that DDT-resistant houseflies detoxified DDT mainly to its noninsecticidal

Chapter | 36  Introduction to Biotransformation (Metabolism)

With glutathione

871

Figure 36.2D  (Continued)

RX + HSCH2CHC(O)NHCH2COOH NHC(O)CH2CH2CH(NH2)COOH glutathione S-transferase RSCH2CHC(O)NHCH2COOH NHC(O)CH2CH2CH(NH2)COOH γ - glutamyltranspeptidase RSCH2CH(O)NHCH2COOH glutamate NH2

cysteinyl glycinase

RSCH2CH(NH2)COOH + glycine N-acetyl transferase RSCH2CHCOOH NHC(O)CH3 Mercapturic acid

R1O R1O

X PSY

X R1O POY R1O

+

GSH

R1O HO R1O

+

GSH

HO R1O R1O

X R1O POY R2

R1O +

GSH

R2 R1O R2

X PSY

+

GSR1

(I)

X POY

+

GSR1

(II)

X POH

+

GSY

(III)

X POH

+

GSY

(IV)

X PSG

+

YOH

(V)

R1=alkyl; R2=aryl; X=S or O; Y=leaving group With thiosulfate

C≡N– Cyanide

+

Na2S2O3=

metabolite DDE. The rate of dehydrohalogenation of DDT to DDE was found to vary between various insect strains as well as between individuals. The enzyme involved, DDTdehydrochlorinase, also occurs in mammals but has been studied more intensively in insects. DDT-dehydrochlorinase, a reduced glutathione (GSH)dependent enzyme, has been isolated from the 100,000 g supernatant of resistant houseflies. Although the enzymemediated reaction requires glutathione, the glutathione concentrations are not altered at the end of the reaction and it is a matter of controversy whether or not this enzyme is a cytosolic glutathione transferase. The lipoprotein enzyme has a molecular mass of 36,000 Da as a monomer and 120,000 Da as the tetramer. The Km for DDT is 5  107M with optimum activity at pH 7.4. This enzyme catalyzes the degradation of p,p-DDT to p,pDDE or the degradation of p,p-DDD (2,2,-bis(p-chloro­ phenyl)-1,1-dichloroethane) to the corresponding DDT

CNS– Thiocyanate

ethylene TDEE (2,2-bis(p-chlorophenyl)-l-chloroethylene). o,p-DDT is not degraded by DDT-dehydrochlorinase, suggesting a p,p-orientation requirement for dehalogenation. In general, the DDT resistance of housefly strains is correlated with the activity of DDT-dehydrochlorinase, although other resistance mechanisms are known in certain strains.

36.2.4  Phase II Reactions: Conjugations Conjugations may be simple, as in the case of phenol, but often they are more complicated processes in which the final product is derived by several steps. In spite of this possible complexity, it is useful to think of conjugation of xenobiotics taking place with glucuronic acid to form glucoronides, N-acetylcysteine to form mercapturic acids, glycine to form hippuric and related acids, sulfate to form ethereal sulfates, thiosulfate ions to form thiocyanate, and glutamine to form conjugates of the same name. In fact, the actual conjugations

Hayes’ Handbook of Pesticide Toxicology

872

COOH With cystine

C≡N– + Cystine Cyanide

COOH

COOH

CH2 CH

CH2 CH

CH2 CH

S . NH2 C

S . NH C

S . N C

N

NH

Figure 36.2e  (Continued)

NH2

2-lminothiocidine-4-carboxylic acid OH With acetate

H3C

OH NH2 +

H3C Acetate

O C CH3 N H

NO2

NO2 2-Amino-6-methyl-4-nitrophenol (a metabolite of DNOC)

2-Acetamido-6-methyl-4-nitrophenol

O H3C S O

Methylation

C2H5O S P S C2H5 Fonofos

3-Hydroxymethyl phenylsulfone

O H3C S O

HS

C2H5O O P S C2H5 Fonofos oxon

OH

O

O H3C S OH O 4-Hydroxymethyl phenylsulfone

often occur with derivatives of the conjugating molecule, for example, with glutathione, uridine diphosphate glu­curonic acid, or phosphoadenine phosphosulfate. Conjugates of foreign chemicals that are rare in mammals, or known only in other classes or phyla, include glucosides, ribosides, ornithines, sulfides, and conjugates with serine, metal complexes, and methylated or acetylated compounds. With the exception of glutathione conjugation, most conjugation reactions involving pesticides are secondary, involving, as substrates, the products of phase I reactions. They include glucoside formation, glucuronic acid formation, sulfate formation, and conjugation with amino acids. This area, as it applies to pesticides, is reviewed in Chapter 38. Conjugation with glutathione, mediated by one of the gluthathione S-transferases, is the first step in a sequence leading to a mercapturic acid (�Figure 36.2). Several pesticides are metabolized in this way, particularly organophosphorus compounds, DDT, -HCH, and organothiocyanates. These reactions and their relationship to pesticides have been reviewed by Motoyama and Dauterman (1980) and by Fukami (1984). The glutathione S-transferases (GSTs) are an abundant family of dimeric proteins that have the capacity to

conjugate glutathione (GSH) with a variety of compounds containing electrophilic centers. The major hepatic cystolic GSTs in mammalian liver can be divided into three classes – alpha (), mu (), pi () – based on sequence similarity and catalytic activity (Mannervik et al., 1985). Each class may contain one or more functional enzymes. Although all of these classes are capable of binding to a wide variety of pesticides, the mu class has somewhat higher affinity than the alpha or pi classes (Dillio et al., 1995; Hayes and Wolf, 1980). Members of the mu class GSTs are responsible for conjugating benzo[a]pyrene7,8-diol-9,10-epoxide (BPDE) as well as a wide variety of pesticides such as the organophosphate insecticides, the halogenated hydrocarbon insecticides, and the S-triazine herbicides (Hayes and Wolf, 1980). Polymorphisms are known to occur in humans in regard to GST enzymes. About 50% of the Caucasian population in the United States is deficient in mu class GSTM1. This polymorphism is due to a deletion in the GSTM1 gene resulting in the lack of GSTM1 protein formation. Epidemiological studies have implicated this deficiency in an increased risk of lung cancer in smokers, presumably due to the ability of GSTM1 to detoxify chemical carcinogens such as BaP in

Chapter | 36  Introduction to Biotransformation (Metabolism)

S

36.3  Major xenobiotic biotransformation reactions

SO 1

FADH-OOH

FADH-OH

NADP+

NADP+

5

O2

2

FADH2

H2O

FAD

NADP+

NADP+

4

3 FAD

NADPH

NADP+

A

O N

C

N

N

R B

FAD

2H•

H

H

O

N

C

N

N

N

R

H

R

FADH2

O2

873

N

OH O

Many of the chemical reactions involved in the biotransformation of pesticides have now been traced to particular enzymes (see Chapter 38), although some are only inferred from the appearance of derivatives of the parent compound in the tissues or excreta of the dosed animal. Chemical reactions reported to occur in the metabolism of pesticides are summarized in Figure 36.2. It should be noted that biotransformation reactions of pesticides may be either detoxications or activations. Hollingworth et al. (1995) provided a detailed review of the detection and significance of active metabolites of pesticides. The biotransformation of most pesticides involves a combination of several chemical reactions and in some instances breakdown products may become part of the general metabolic pool. For example, formaldehyde formed in demethylation reactions may be incorporated into the one-carbon metabolic pool.

O C N

hydroperoxyflavoprotein

Figure 36.3  Catalytic cycle for flavin-containing monooxygenase (FMO)-catalyzed monooxygenase reactions.

tobacco smoke (Bell et al., 1992; Nakachi et al., 1993; Seidegard and Pero, 1985; Wormhoudt et al., 1999). Because conjugation reactions other than those mediated by the GSTs are less well known in the metabolism of pesticides, the enzymatic basis of these conjugations is not discussed in detail. However, this matter has been reviewed in detail by Motoyama and Dauterman (1980), Dorough (1984), Matsumura (1985), and Hollingworth et al. (1995) and several types of conjugation are known to involve pesticides. Glucuronides are important in the metabolism of carbamates such as banol, carbaryl, and carbofuran (Mehendale and Dorough, 1972) as well as some organophosphate compounds (Hutson, 1981) and other chemicals. Ethereal sulfates, while less important in the metabolism of pesticides than glucuronides, nevertheless may be formed from carbofuran and other carbamates (Dorough, 1968). Glutathione conjugation is important in the metabolism of organophosphates (Motoyama and Dauterman, 1980) and the conjugated products of glutathione adducts may be further metabolized to mercapturic acids, the N-acetyl cys­ teine derivative of the original xenobiotic substrate. Insects and plants are unusual in forming glucosides rather than glucuronides. A general review of phase II metabolism of xenobiotics is that of LeBlanc (2008).

Conclusions Pesticides are subject to modification by a wide range of phase I and phase II enzymes and their isoforms and polymorphic variants. The products are numerous and secondary modification of the primary metabolites complicates the situation even further. Since the products may be more (activation) or less (detoxication) toxic than the parent chemical, knowledge of metabolism may be critical in the extrapolation from experimental animal to humans necessary for accurate human health risk assessment.

References Adams, N. H., Levi, P. E., and Hodgson, E. (1990). In vitro studies of the metabolism of atrazine, simazine and terbutryn in several vertebrate species. J. Agric. Food Chem. 28, 1411–1417. Arand, M., Cronin, A., Adamska, M., and Oesch, F. (2005). Epoxide hydrolases: structure, function, metabolism and assay. Methods Enzymol. 400, 569–588. Bell, D. A., Thompson, C. L., Taylor, J., Miller, C. R., Perera, F., Hsieh, L. L., and Lucier, G. W. (1992). Genetic monitoring of human polymorphic cancer susceptibility genes by polymerase chain reaction: Application to glutathione transferase . Environ. Health Perspect. 98, 113. Bogaards, J. J. P., van Ommen, B., Wolf, C. R., and van Bladeren, P. J. (1995). Human cytochrome P450 enzyme selectivities in the oxidation of chlorinated benzenes. Toxicol. Appl. Pharmacol. 132, 44–52. Butler, A. M., and Murray, M. (1997). Biotransformation of parathion in human liver: Participation of CYP3A4 and its inactivation during microsomal parathion oxidation. J. Pharmacol. Exp. Ther. 280, 966–973. Chambers, J. E., and Carr, R. L. (1995). Biochemical mechanisms contributing to species differences in insecticidal toxicity. Toxicology 105, 291–304.

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Cherrington, N. J., Can, Y., Cherrington, J. W., Rose, R. L., and Hodgson, E. (1998a). Physiological factors affecting protein expression of flavincontaining monooxygenases 1, 3 and 5. Xenobiotica 7, 673–682. Cherrington, N. J., Falls, J. G., Rose, R. L., Clements, K. M., Philpot, R. M., Levi, P. E., and Hodgson, E. (1998b). Molecular cloning, sequence, and expression of mouse flavin-containing monooxygenases 1 and 5 (FMO1 and FMO5). J. Biochem. Mol. Toxicol. 12, 205–212. Coleman, S., Liu, S., Linderman, R., Hodgson, E., and Rose, R. L. (1999). In vitro metabolism of alachlor by human liver microsomes and human cytochrome P450 isoforms. Chem.-Biol. Interact 122, 27–39. Coleman, S., Linderman, R., Hodgson, E., and Rose, R. L. (2000). Comparative metabolism of chloracetamide herbicides and selected metabolites in human and rat liver microsomes. Environ. Health Perspect. 108, 1151–1157. Dillio, C., Sacchetta, P., Jannarelli, V., and Aceto, A. (1995). Binding of pesticides to alpha class, mu-class and pi-class glutathione transferases. Toxicol. Lett. 76, 173–177. Dorough, H. W. (1968). Metabolism of Furadan (NIH-10242) in rats and house-flies. J. Agric. Food Chem. 16, 319–325. Dorough, H. W. (1984). Metabolism of insecticides by conjugation reactions. In “Differential Toxicities of Insecticides and Halogenated Aromatics” (F. Matsumura, ed.), Int. Encycl. Pharmacol. Ther., Sect. 113, pp. 291–330. Pergamon, New York. Dorough, H. W., and Casida, J. E. (1964). Nature of certain carbamate metabolites of the insecticide Sevin. J. Agric. Food Chem. 12, 294–304. Ecobichon, D. J. (2001). Toxic effects of pesticides. In “Casarett and Doull’s Toxicology, the Basic Science of Poisons” (C. D. Klaassen, ed.), 6th ed. McGraw-Hill, New York. Eling, T., Boyd, J., Reed, G., Mason, R., and Sivarajoh, K. (1983). Xenobiotic metabolism by prostaglandin endoperoxide synthetase. Drug Metab. Rev. 14, 1023. Fukami, J. (1984). Metabolism of several insecticides by glutathione S-transferase. In “Differential Toxicities of Insecticides and Halogenated Aromatics” (F. Matsumura, ed.), Int. Encycl. Pharmacol. Ther., Sect. 113, pp. 223–264. Pergamon, New York. Gonzalez, F. J. (1990). Molecular genetics of the P450 superfamily. Pharmacol. Ther. 45, 1–38. Guengerich, F. P., and MacDonald, T. L. (1984). Chemical mechanisms of catalysis by cytochromes P450: A unified view. Acc. Chem. Res. 17, 9–16. Hajjar, N. P., and Hodgson, E. (1980). Flavin adenine dinucleotide-dependent monooxygenase: Its role in the sulfoxidation of pesticides in mammals. Science 209, 1134–1136. Hajjar, N. P., and Hodgson, E. (1982a). Sulfoxidation of thioether-containing pesticides by the flavin-adenine dinucleotide-dependent monooxygenase of pig liver microsomes. Biochem. Pharmacol. 31, 745–752. Hajjar, N. P., and Hodgson, E. (1982b). The microsomal FAD-dependent monooxygenase as an activating enzyme: Fonofos metabolism. In “Biological Reactive Intermediates” (R. Synder, D. U. Porke, J. J. Kocsis, D. Jollow, G. G. Gibson, and C. Witmer, eds.), Vol. 2, Part B, pp. 1245–1253. Plenum, New York. Hanioka, N., Jinno, H., Toshiko, T.-K., Nishimura, T., and Ando, M. (1999). In vitro metabolism of chlorotriazines: Characterization of simazine, atrazine, and propazine metabolism using liver microsomes from rats treated with various cytochrome P450 inducers. Toxicol. Appl. Pharmacol. 156, 195–205. Hayes, J. D., and Wolf, C. R. (1980). Role of glutathione in drug resistance. In “Glutathione Conjugation: Its Mechanisms and Biological

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Significance” (H. Sies and B. Ketterer, eds.), pp. 315–355. Academic Press, London. Hodgson, E. (1974). Comparative studies of cytochrome P450 and its interaction with pesticides. In “Survival in Toxic Environments” (M. A. Q. Khan and J. P. Bederka Jr., eds.), pp. 213–260. Academic Press, New York. Hodgson, E. (1982–1983). Production of pesticide metabolites by oxidative reactions. J. Toxicol. Clin. Toxicol. 19, 609–621. Hodgson, E., and Kulkarni, A. P. (1974). Interactions of pesticides with cytochrome P450. ACS Symp. Ser. 2, 14–38. Hodgson, E., and Levi, P. E. (eds) (1994). “Introduction to Biochemical Toxicology,” 2nd ed. Appleton & Lange, Stamford, CT. Hodgson, E., and Levi, P. E. (1997). “A Textbook of Modern Toxicology,” 2nd ed. Appleton & Lange, Stamford, CT. Hodgson, E., and Meyer, S. A. (1997). Pesticides. In “Hepatic and Gastrointestinal Toxicology” (R. S. McCluskey and D. L. Earnest, eds.), Comprehensive Toxicology (I. G. Sipes, C. A. McQueen, and A. J. Gandolfi, series eds.). Pergamon, Oxford, UK. Hodgson, E., Rose, R. L., Ryu, D.-Y., Falls, G., Blake, B. L., and Levi, P. E. (1995). Pesticide-metabolizing enzymes. Toxicol. Lett. 82–83, 73–81. Hodgson, E., Cherrington, N., Coleman, S. C., Liu, S., Falls, J. G., Cao, Y., Goldstein, J. E., and Rose, R. L. (1998). Flavin-containing monooxygenase and cytochrome P450 mediated metabolism of pesticides: From mouse to human. Rev. Toxicol. 2, 231–243. Hodgson, E., Das, P. C., Cho, T. M., and Rose, R. L. (2008). Phase I metabolism of toxicants and metabolic interactions. In “Molecular and Biochemical Toxicology” (R. C. Smart and E. Hodgson, eds.). John Wiley and Sons, Hoboken NJ. Hollingworth, R. M., Kurihara, N., Miyamoto, J., Otto, S., and Paulson, G. D. (1995). Detection and significance of active metabolites of agrochemicals and related xenobiotics in animals. Pure Appl. Chem. 67, 1487–1532. Hutson, D. H. (1981). The metabolism of insecticides in man. Prog. Pestic. Biochem. 1, 287–333. Inui, H., Kodama, T., Ohkawa, Y., and Ohkawa, H. (2000). Herbicide metabolism and cross-tolerance in transgenic potato plants co-expressing human CYP1A1, CYP2B6, and CYP2C19. Pestic. Biochem. Physiol. 66, 116–129. Jakoby, W. B., ed. (1980). “Enzymatic Basis of Detoxication”, Vols. 1 and 2. Academic Press, New York. Jakoby, W. B., Bend, J. R., and Caldwell, J. (eds) (1982). “Metabolic Basis of Detoxication.” Academic Press, New York. Kinsler, S., Levi, P. E., and Hodgson, E. (1990). Relative contributions of cytochrome P450 and flavin-containing monooxygenases to the microsomal oxidation of phorate following treatment of mice with phenobarbital, hydrocortisone, acetone, and piperonyl butoxide. Pestic. Biochem Physiol. 37, 174–181. Klaassen, C. D., ed. (2001). “Casarett and Doull’s Toxicology,” 6th ed. McGraw-Hill, New York. Kulkarni, A. P., and Hodgson, E. (1980). Metabolism of insecticides by the microsomal mixed function oxidase system. Pharmacol. Ther. 8, 397–475. Kulkarni, A. P., and Hodgson, E. (1984a). Metabolism of insecticides by the microsomal mixed function oxidase systems. In “Differential Toxicities of Insecticides and Halogenated Aromatics” (F. Matsumura, ed.), Int. Encycl. Pharmacol. Ther., Sect. 113, pp. 27–128. Pergamon, New York. Kulkarni, A. P., and Hodgson, E. (1984b). The metabolism of insecticides: The role of monooxygenase enzymes. Annu. Rev. Pharmacol. Toxicol. 24, 19–42.

Chapter | 36  Introduction to Biotransformation (Metabolism)

Lang, D., Criegee, D., Grothusen, A., Saalfrank, R. W., and Bocker, R. H. (1996). In vitro metabolism of atrazine, terbutylazine, ametryne, and terbutryne in rats, pigs and humans. Drug Metab. Dispos. 24, 859–865. Lang, D. H., Rettie, A. E., and Bocker, R. H. (1997). Identification of enzymes involved in the metabolism of atrazine, terbutylazine, amet­ ryne, and terbutryne in human liver microsomes. Chem. Res. Toxicol. 10, 1037–1044. Lawton, M. P., and Philpot, R. M. (1995). Emergence of the flavincontaining monooxygenase gene family. Rev. Biochem. Toxicol. 11, 1–27. Lawton, M. P., Cashman, J. R., Cresteil, T., Dolphin, E. T., Elfarra, A. A., Hines, R. N., Hodgson, E., Kimura, T., Ozols, J., Phillips, I. R., Philpot, R. M., Poulsen, L. L., Rettie, A. E., Shephard, E. A., Williams, D. E., and Ziegler, D. M. (1994). A nomenclature for the mammalian flavincontaining monooxygenase gene family based on amino acid sequence identities. Arch. Biochem. Biophys. 308, 254–357. LeBlanc, G. A. (2008). Phase I – Conjugation of toxicants. In “Molecular and Biochemical Toxicology” (R. C. Smart and E. Hodgson, eds.). John Wiley and Sons, Hoboken NJ. Levi, P. E., and Hodgson, E. (1984). Oxidation of pesticides by purified cytochrome P450 isozymes from mouse liver. Toxicol. Lett. 24, 221–228. Levi, P. E., and Hodgson, E. (1988). Stereospecificity of the oxidation of phorate and phorate sulphoxide by purified FAD-containing monooxygenase and cytochrome P450. Xenobiotica 1, 29–39. Levi, P. E., and Hodgson, E. (1992). Metabolism of organophosphorus compounds by the flavin-containing monooxygenase. In “Organophosphates, Chemistry, Fate, and Effects” (J. E. Chambers and P. E. Levi, eds.), pp. 141–154. Academic Press, New York. Magdalou, J., and Hammock, B. D. (1987). Metabolism of tridiphane (2-(3,5-dichlorophenyl)-2-(2,2,2-trichloroethyl)oxirane) by hepatic epoxide hydrolases and glutathione S-transferases in mouse. Toxicol. Appl. Pharmacol. 91, 439–449. Mailman, R. B., and Hodgson, E. (1972). The cytochrome P450 substrate optical difference spectra of pesticides with mouse hepatic microsomes. Bull. Environ. Contam. Toxicol. 8, 186–192. Mailman, R. B., Kulkarni, A. P., Baker, R. C., and Hodgson, E. (1974). Cytochrome P450 difference spectra: Effect of chemical structure on type II spectra in mouse hepatic microsomes. Drug Metab. Dispos. 2, 301–311. Mannervik, B., Alin, P., Guthenberg, C., Jennson, H., Tahir, M. K., Warhom, M., and Jornvall, H. (1985). Identification of 3 classes of cytosolic glutathione transferase common to several mammalian species: Correlation between structural data and enzymatic properties. Proc. Natl. Acad. Sci. USA 82, 7202. Marnett, L. J., and Eling, T. E. (1983). Cooxidation during prostaglandin biosynthesis: A pathway for the metabolic activation of xenobiotics. Rev. Biochem. Toxicol. 5, 135–172. Matsumura, F. (1985). “Toxicology of Insecticides,” 2nd ed. Plenum, New York. Mehendale, H. M., and Dorough, H. W. (1972). Conjugative metabolism and action of carbamate insecticides. In “Insecticide-Pesticide

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Chemistry” (A. S. Tahori, ed.), pp. 37–49. Gordon & Breach, London. Motoyama, N., and Dauterman, W. C. (1980). Glutathione S-transferases: Their role in the metabolism of organophosphorus insecticides. Rev. Biochem. Toxicol. 2, 49–70. Nakachi, K., Inai, K., Hayashi, S., and Kawajiri, K. (1993). Polymorphisms of the CYP1A1 and glutathione S-transferase genes associated with susceptibility to lung cancer in relation to cigarette dose in a Japanese population. Cancer Res. 53, 2994. Nebert, D. W., Nelson, D. R., and Feyereisen, R. (1989). Evolution of the cytochrome P450 genes. Xenobiotica 19, 1149–1160. Nelson, D. R. (2008) (http://drnelson.utmem.edu/p450stats.Feb2008.htm). Rose, R. L., Hodgson, E., and Roe, R. M. (1999). Pesticides. In “Toxicology” (H. Marquardt, S. G. Schafer, R. O. McClellan, and F. Welsch, eds.). Academic Press, San Diego. Satoh, T. (1987). Role of carboxylases in xenobiotic metabolism. Rev. Biochem. Toxicol. 8, 155–182. Seidegard, J., and Pero, R. W. (1985). The hereditary transmission of high glutathione transferase activity toward trans-stilbene oxide in human mononuclear leukocytes. Hum. Genet. 69, 66. Smart, R. C., and Hodgson, E. (2008). “Molecular and Biochemical Toxicology.” John Wiley and Sons, Hoboken NJ. Smyser, B. P., and Hodgson, E. (1985). Metabolism of phosphorus-containing compounds by pig liver microsomal FAD-containing monooxygenase. Biochem. Pharmacol. 34, 1145–1150. Smyser, B. P., Sabourin, P. J., and Hodgson, E. (1985). Oxidation of pesticides by purified microsomal FAD-containing monooxygenase from mouse and pig liver. Pestic. Biochem. Physiol. 24, 368–374. Tang, J., Rose, R. L., Brimfield, A. A., Dai, D., Goldstein, J. A., and Hodgson, E. (2001). Metabolism of chlorpyrifos by human cytochrome P450 isoforms and human, mouse and rat liver microsomes. Drug Metabol. Disp. 29, 1201–1204. Tynes, R. E., and Hodgson, E. (1985). Magnitude of involvement of the mammalian flavin-containing monooxygenase in the microsomal oxidation of pesticides. J. Agric. Food Chem. 33, 471–479. Wilkinson, C. F., ed. (1976). “Insect Biochemistry and Physiology”. Plenum, New York. Williams, R. T. (1959). Detoxication mechanisms. In “The Metabolism and Detoxication of Drugs, Toxic Substances and Other Organic Compounds” (R. T. Williams, ed.), 2nd ed. Wiley, New York. Wormhoudt, L. W., Cammandeur, J. N. M., and Vermeulen, N. P. E. (1999). Genetic polymorphisms of human N-acetyltransferase, cytochrome P450, glutathione-S-ansferase, and epoxide hydrolase enzymes: Relevance to xenobiotic metabolism and toxicity. Crit. Rev. Toxicol. 29, 59–124. Zeldin, D. C., and Seubert, J. M. (2008). Structure, mechanism and regulation of cytochromes P450. In “Molecular and Biochemical Toxicology” (R. C. Smart and E. Hodgson, eds.). John Wiley and Sons, Hoboken NJ. Ziegler, D. M. (1980). Microsomal flavin-containing monooxygenase: Oxygenation of nucleophilic nitrogen and sulfur compounds. In “Enzymatic Basis of Detoxication” (W. B. Jacoby, ed.), Vol. 1, Chap. 9. Academic Press, New York.

Chapter 37

Absorption Ronald E. Baynes and Jim E. Riviere North Carolina State University, Raleigh, North Carolina

37.1  Introduction For a pesticide to elicit toxicity, it must be transferred from the external site of exposure to the target site (e.g., organ, nucleic acid, receptor) and achieve a sufficiently high con­ centration in the target organ (Figure 37.1). Absorption is the translocation of the pesticide from an external source of exposure to the bloodstream. Once in the blood, the chemi­ cal is distributed through the body and delivered to tissues, where it may leave the blood and enter the cells of the tis­ sue or it may remain in the blood and simply pass through the tissue. In certain tissues such as the liver, the chemical may be effectively removed from the body by metabolism.

Oral

Dermal

Gastrointestinal Tract

Skin

Other tissues, such as kidney and lung, serve to eliminate xenobiotics from the body by excretion. Absorption, distri­ bution, metabolism, and excretion, which are collectively termed disposition, are all factors that affect the concen­ tration of a chemical in target tissues. Pharmacokinetics refers to the mathematical description of the time course of chemical disposition in the body. Metabolism and excretion are discussed in detail in other chapters of this work. This chapter focuses on pesticide absorption with an expanded focus on dermal absorption as workers involved in pesticide manufacturing, formulation, or application and harvesting of treated crops, are more likely to be exposed to these chemicals via the skin.

iv, ip, sc, im

Inhalation

Lung

Excretion

Excretion Liver

Blood

Metabolites Kidney

Excretion Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

Target Organ(s) Figure 37.1  Representation of the absorption, distribution, metabol­ ism, and excretion of toxicants.

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37.2  Factors that influence the transfer and availability of chemicals in the body

membrane is proportional to the concentration gradient, the membrane surface area, and the permeability coefficient of the molecule. The permeability coefficient is the product of the partition coefficient and the diffusion coefficient.

For the routes of pesticide exposure relevant to humans, the pesticide must cross one or more cell membranes to reach the bloodstream, then one or more additional cell mem­ branes to leave the blood and enter tissues. The following discussion concerns the factors that influence the transfer of chemicals across biological membranes.

37.2.1  Properties of Cell Membranes Cell membranes (i.e., plasma membranes) consist of phospholipids and proteins (Figure 37.2). The fluid and dynamic phospholipid bilayer, with polar head groups on the intracellular and extracellular surfaces and fatty acid chains filling the inner space, acts as a permeability barrier to water-soluble molecules. Proteins interspersed through­ out the phospholipid bilayer mediate the transport of small water-soluble molecules into and out of the cell by forming pores or by acting as carriers. Molecules cross membranes by passive transport, which requires the expenditure of no energy, or by specialized transport systems. The ability of a chemical to cross various membrane barriers is determined by its physicochemical properties, which include lipophi­ licity, molecular size, and ionization.

37.2.2  Transport Mechanisms 37.2.2.1  Passive Transport Passive transport occurs by simple diffusion or via pores in the plasma membrane (Figure 37.2). Most lipophilic molecules cross membranes by simple diffusion in accord with Fick’s first law of diffusion (see equation below), which states that the flux or rate at which a molecule diffuses across the plasma



diffusion coefficient  surface area  partition coefficient  conc. gradient flux  skin or membrane thickness

Theoretically, the determinants of flux or diffusion rates across the skin or gastrointestinal (GI) tract may be altered clinically or experimentally through manipulation of pesticide formulations. If lipid solubility increases, the penetrant may remain in the stratum corneum of the skin and form a reservoir. Some compounds can also form a reservoir in the dermis. These scenarios can prolong absorption half-life across the skin, which can also prolong the body burden of the penetrant. Ingestion of very lipid soluble pesticides, which are not miscible in the aqueous intestinal fluid, can be presented as emulsions and brought into solution through the action of detergent-like bile acids. The product of this mixing is large surface area micelles (hydrophobic interior) that deliver the lipids to the brush border of the intestine for diffusion across the membrane. Water readily traverses the plasma membrane through pores and may carry with it small hydrophilic solutes. The pores in most cells are approximately 4 Å in diameter. In the kidney glomeruli, however, the pores may be as large as 70–80 Å in diameter, which permits more efficient renal elimination of potentially toxic compounds. Weak organic acids and bases may cross plasma membranes by simple diffusion when they are nonionized. Ionized weak organic acids and bases, how­ ever, slowly permeate the plasma membrane through pores. According to the Brønsted–Lowry theory, an acid is a proton donor and a base is a proton acceptor. The ratio of nonionized to ionized molecules of a weak organic acid or base depends on the dissociation constant (Ka) and the pH of the media (Table 37.1). The dissociation constant is

Extracellular Space

Phospholipid Bilayer

Intracellular Space Energy

Simple Diffusion

Diffusion via Pores

Active Transport

Figure 37.2  Schematic of the plasma membrane and mechanisms of transport across the membrane.

Facilitated Diffusion

Chapter | 37  Absorption

879

usually expressed in terms of its negative logarithm, and the relationship between pKa and pH is derived from the Henderson–Hasselbalch equation as shown in Table 37.1. The pKa is the pH at which 50% of the acid or base is ion­ ized. The concept of pKa is particularly important for oral absorption (see Section 37.3.2) and often overlooked when assessing dermal absorption (see Section 37.3.1). The penetration of acidic and basic pesticides through skin can be influenced by the skin surface pH, which is weakly acidic (pH 4.2–5.6). Paraquat and diquat are hydro­ philic pesticides that exist as fixed charged cations and remain dissociated at all pH values. Very little paraquat or diquat is, therefore, expected to be absorbed by skin, although percutaneous absorption of paraquat has resulted in systemic effects and deaths in humans (Smith, 1988). Dermal absorption studies in human volunteers demon­ strated 0.29, 0.23, and 0.29% absorption in the leg, forearm, and forearm, respectively (Wester et al., 1984). Other stud­ ies have determined that the in vitro permeability constants for paraquat in various animal species (rat, hairless rat, nude rat, mouse, hairless mouse, rabbit, guinea pig) are 40–1600 times greater than for humans (Walker et al., 1983). One radiolabeled in vivo rat study reported a dermal bioavail­ ability of 3.8% (Chui et al., 1988), which supports the claim that rodent studies can overestimate human absorption. Like paraquat, very little diquat is absorbed (0.3%) in the human forearm in vivo (Maibach and Feldmann, 1974). Diquat absorption increased to 1.4% with occlusion and to 3.8% with damaged skin. Data from these in vivo and in vitro studies suggest that paraquat- or diquat-induced dermatotox­ icity is a highly probable mechanism, a priori, for dermal absorption of these hydrophilic and charged pesticides.

Facilitated diffusion is similar to active transport, except that the solute moves only in the direction of a concentration or electrochemical gradient and the expenditure of energy is not required (Figure 37.2). Additional types of special­ ized transport are exocytosis and endocytosis, processes by which cells secrete and ingest large molecules, respectively. There are two types of endocytosis; pinocytosis (cell drink­ ing), which is the ingestion of fluids and solutes, and phago­ cytosis (cell eating), which is the ingestion of large particles. Phagocytosis is especially important in the removal of par­ ticulate matter in the respiratory tract. Recent studies have also suggested an even finer gradation in specific transport processes (e.g., calveolae) that facilitate entry of differentsized material into the cell. Many of the available commercial pesticides are trans­ ported across the skin and GI tract by passive diffusion. However, there is some evidence that membrane transport pro­ teins play a significant role in the absorption mechanism in the GI tract and accounts for pesticide influx and/or efflux of sev­ eral pesticides. The hydrophilic herbicide, paraquat, is thought to be absorbed by a mechanism that consists of facilitated, saturable, and diffusional components (Heylings, 1991; Nagao et al., 1993). The P-glycoprotein (P-gp/MDR1), is a transmem­ brane transporter in humans and animals that is encoded by the ABCB1/MDR1 gene. This transporter is in various human tis­ sues such as the apical surface of intestinal epithelial cells. The interaction between P-glycoproteins and the avermectin class pesticides and other classes of insecticides such as methylpara­ thion, endosulfan, cypermethrin, and fenvalerate have been well documented (Sreeramulu et al., 2007; Zhou, 2008). A similar or related mechanism has been reported for the influx and efflux of neonicotinoids (Brunet et al., 2008). These inter­ actions are important as they dictate the rate and extent of pes­ ticide absorption across the intestine, especially when there is co-exposure to other drugs or pesticides that may compete with the pesticide of interest and consequently increase pesti­ cide uptake across the GI tract (Alvinerie et al., 2008).

37.2.2.2  Specialized Transport Active transport systems are characterized by (1) movement of solutes against a concentration or electrochemical gradi­ ent, (2) saturation at high solute concentration, (3) specific­ ity for structural and/or chemical features of the solute, (4) competitive inhibition by molecules transported by the same transporter, and (5) inhibition of transport by compounds and/or processes that interfere with cellular metabolism.

37.2.3  Protein (Macromolecular) Binding Blood consists of red blood cells, white blood cells, and platelets suspended in plasma. Plasma, which comprises

Table 37.1  Acids and Bases According to the Brønsted-Lowry Theory Acid

Base

Representation

AH ↔ A  H

B  H ↔ BH

Definition

Proton donor (AH)

Proton acceptor (B)

Dissociation constant (Ka)

1 pKa  log K    log Ka a



Ka 



[A ][H ] [AH]

pKa  pH  log

Ka  [AH] [A ]

[B][H ] [BH ]

pKa  pH  log

[BH ] [B]

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approximately 55% of the blood volume in humans, also contains a number of proteins, ions, and inorganic mole­ cules. Many xenobiotics in blood are reversibly bound to plasma proteins, including albumin, 1-acid glycoprotein, lipoproteins, and globulins. Reversible binding to plasma proteins enhances the solubility of lipophilic compounds in blood and influences the rate of distribution to tissues. Proteins are amphoteric in nature and therefore possess cat­ ionic and anionic regions. Many acidic chemicals bind to albumin, whereas basic chemicals tend to bind to 1-acid glycoprotein and lipoproteins. The high molecular weight of proteins prevents them, and any toxicants they bind, from crossing cell membranes. Only the free (or unbound) chemi­ cal is available to cross plasma membranes (Figure 37.3). The interaction of chemicals and plasma proteins, however, is rapid and reversible. Equilibrium is quickly established between the bound and the unbound forms of the chemi­ cal. As unbound chemical crosses a plasma membrane in a microenvironment, bound chemical dissociates to reestab­ lish equilibrium with the unbound fraction. Gomez-Catalan et al. (1991) investigated the distribu­ tion of various organochlorines in rat and human blood. In rat blood, 87% of hexachlorobenzene was associated with red blood cells, approximately 84% of DDE was bound to plasma proteins, lindane was nearly equally dis­ tributed between red blood cells and plasma, and 97% of pentachlorophenol was associated with plasma. In plasma, lindane (64%) and DDE (92%) were mainly associated with lipoproteins, pentachlorophenol was mainly associ­ ated with “other” plasma proteins (81%), and hexachlorobenzene was nearly equally distributed. A very differ­ ent pattern of distribution was observed in human blood. Hexachlorobenzene and lindane in plasma were nearly equally distributed between lipoproteins and “other” plasma proteins, whereas 60% of DDE was associated with “other” proteins. Other investigators have shown that

Blood

Tissue

Bound

Bound

Free

Free

Figure 37.3  Equilibrium is established between free (unbound) and bound xenobiotic in blood and between free xenobiotic in blood and tis­ sues. Only the free xenobiotic crosses the plasma membrane, which is rep­ resented by the dashed line separating the blood and tissue compartments.

dieldrin is greater than 99% bound to human serum pro­ teins (Garrettson and Curley, 1969), and diflubenzuron is 40–50% bound to plasma proteins in chickens (Opdycke and Menzer, 1984). The organophosphate diazinon is 89% bound to proteins in rat plasma (Wu et al., 1996).

37.3  Absorption Human exposure to pesticides is typically by oral, dermal, and inhalation routes. Occupational exposure to pesticides is more likely to occur by dermal contact, and this will be the focus of Section 37.3.1 of the chapter. Percutaneous absorption is reported as the possible route of entry in 65–85% of all cases of occupational exposure with pes­ ticides (Galli and Marinovich, 1987). Spray or dusting of pesticides can result in disposition of 20–1700 times the amount deposited in the respiratory tract (Feldmann and Maibach, 1974). Epidemics of pesticide poisoning following cutaneous exposure have been reported for nonoccupational uses (Ferrer and Cabral, 1993). These cases often involved accidental contamination of infant clothing or exposure to talcum powder with pesticides (Martin-Bouyer et al., 1983). These anecdotal case reports, coupled with dermal exposure estimated from various direct and indirect dosimetric exper­ iments are often the only available human data with which to perform dermal absorption assessment. In spite of such limited data, it is possible to estimate dermal absorption by extrapolation from dermal exposure data. Algebraic equa­ tions that take into account exposure time and the chemi­ cal nature of the compound (lipophilicity and molecular weight) have been presented for estimating dermal absorp­ tion (Cleek and Bunge, 1993; Potts and Guy, 1992). Exposure of small children is more likely to be by oral and/or dermal routes. Absorption from the gastrointestinal and respiratory tracts is discussed in Sections 37.3.2 and 37.3.3, respectively. Other routes of exposure that are used primarily in the laboratory (subcutaneous, intravenous, intraperitoneal, and intramuscular) are discussed only briefly in Section 37.3.4 A chemical is considered to be absorbed when it reaches the bloodstream. For routes other than intravenous administration, which bypasses the process of absorption, a chemical is absorbed when it crosses the epithelial lay­ ers in the skin, small intestine, or alveoli in the lungs and enters the bloodstream from an external site of exposure. Compared to the epithelium in the small intestine, the skin is relatively impermeable to aqueous solutions and ions, but it may be permeable in varying degrees to a large number of drugs or xenobiotics. Drug or xenobiotic delivery path­ ways in the skin and GI tract can hypothetically involve intercellular and intracellular passive diffusion across the epidermis and transappendageal routes via hair follicles and sweat pores in the skin. Transappendageal pathways are considered to contribute very little to the dermal transport

Chapter | 37  Absorption

of most drugs compared to transport across the epidermis (Barry, 1991). It is possible for very small and/or polar molecules to penetrate through these appendages or shunts, but very unlikely for many classes of highly lipophilic pesticides. The stratum corneum cell layer in human skin (10–50 m) and pig skin (15 m) is nonviable and is consid­ ered to be the rate-limiting barrier in percutaneous absorp­ tion of many drugs and pesticides (Monteiro-Riviere et al., 1990). Most available research has concentrated on the stra­ tum corneum as the primary barrier to absorption, although the viable epidermis (ca. 80 m in humans and 60 m in pigs) and dermis (3–5 mm in humans) may contribute sig­ nificantly to the percutaneous penetration of drugs and ulti­ mately their bioavailability. Scheuplein (1972) proposed that polar drugs diffused through the hydrated keratin of the dead cells in the stratum corneum, whereas nonpolar drugs traversed the intracellular lipid. The accepted hypothesis is that the dominant pathway for polar molecules resides in the aqueous region of the intercellular lipid with the hydro­ phobic region of the lipid chains providing the nonpolar route (Elias, 1981). The intercellular region as depicted in the brick and mortar model of the stratum corneum, now considered the most likely path for absorption of lipophilic drugs and pesticides, is filled with neutral lipids (complex hydrocarbons, free sterols, sterol esters, free fatty acids, and triglycerides), which makes up 75% of the total lipids, and polar lipids, such as phosphatidylethanolamine, phos­ phatidylcholine, lysolecithin, ceramides, and glycolipids (Magee, 1991). Percutaneous and GI absorption through the intercellular pathway is by passive diffusion and it is often correlated to the partition coefficient. The rate of absorption of the penetrant can be described by Fick’s law of diffusion. Continuous blood flow removes the xenobiotic from the site of absorption in the skin and GI tract, thus main­ taining a concentration gradient and enhancing continued absorption. For many of the lipophilic pesticides, penetrat­ ing molecules are thought to enter the systemic circulation at the dermis/epidermis interface in skin and do not nec­ essarily traverse the full thickness of the dermis. For rap­ idly absorbed chemicals, equilibrium may be established between the blood and the site of absorption, and the rate of entry into the blood is limited by blood flow rather than by diffusion across the membrane. In this case, an increase in blood flow will increase the rate of absorption of the chemical and absorption is said to be perfusion- (or blood-flow) limited. For poorly absorbed chemicals, how­ ever, absorption is not sensitive to blood flow and is said to be diffusion-rate-limited.

37.3.1  Percutaneous Absorption The skin is a complex tissue with a large surface area whose primary function is to protect the body from physi­ cal or chemical insult, to thermoregulate and to simultane­ ously prevent water loss from the body. Dermal absorption

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of any chemical requires movement from the environment across this barrier, which is a biochemical milieu of com­ plex lipids and proteins. Experimentally, there are several in vitro, ex vivo, and in vivo models that have been used to esti­ mate dermal absorption of pesticides in humans. Although in vivo methods are the gold standard, each of these methods has its respective weaknesses and strengths for accurately predicting dermal absorption of pesticides. Dermal absorp­ tion assessment is further complicated by species, age, and sex differences, and differences between anatomical sites within a species. More importantly, dermal absorption in rodent skin is not always equivalent to that in human skin. Dermal absorption is dependent on the physicochemical properties of the pesticide, the formulation, and the environ­ mental conditions. The pesticide applicator is often clothed and operating in extreme environments, not standard labo­ ratory conditions. This section of the chapter will focus on the differences in absorption between anatomical body sites (Section 37.3.1.1) and the effects of formulation chemis­ try (Section 37.3.1.2) and environmental factors (Section 37.3.1.3) that influence percutaneous absorption of pesticides.

37.3.1.1  Anatomical Site Differences Regional variation in skin permeability in different body sites may be related to skin thickness, number of cell lay­ ers, cell size of the epidermis and stratum corneum, and distribution of hair follicles and sweat pores. Because of thick layers of stratum corneum, permeability in palmar and plantar skin is expected to be less than that in the scalp or forearm (Feldmann and Maibach, 1974). Data from several studies suggest that regional variation in vas­ cular anatomy and blood flow should also be considered (Monteiro-Riviere et al., 1990; Qiao et al., 1993). Various studies have demonstrated regional variation in penetration of drugs and pesticides in pig skin (Qiao and Riviere, 1995; Qiao et al., 1993), rat skin (Bronaugh, 1985), and rhesus monkey skin (Wester et al., 1980). These studies further demonstrated that parathion penetrates non­ occluded pig skin in the decreasing order back  shoul­ der  buttocks  abdomen; for occluded skin, the order is back  abdomen  buttocks  shoulder. Wester et al. (1994) also demonstrated that pyrethrin absorption through the human forearm is less than the predicted absorption in the human scalp. This anatomical difference is somewhat consistent with lindane absorption through the forearm (18%), forehead (34%), and palm (34%) of rhesus mon­ keys (Moody and Ritter, 1989). This anatomical range for lindane is similar to that for dermal absorption of DEET (diethyl-m-toluamide) in rhesus monkeys (Moody et al., 1989). There are also significant data to suggest that der­ mal absorption of permethrin, aminocarb, DEET, and fenitrothion in monkey foreheads is twice that in monkey forearms (Moody and Franklin, 1987; Moody et al., 1987; Sidon et al., 1988).

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However, Moody et al. (1990, 1992) demonstrated that there is no difference between the absorption of acid and amine forms of 2,4-D in rhesus monkey forearm and fore­ head and forearm and palm regions. The palmar absorp­ tion data conflict with the accepted dogma that absorption through palmar skin should theoretically be less than that in forearm skin because of the thickness of the stratum corneum in palmar skin (Maibach et al., 1971). It is proposed that because of the hydrophilic nature of 2,4-D-amine, absorption can occur through polar routes such as eccrine glands, which are more frequent in the palmar skin than in the forearm skin. This anatomical difference does not explain the discrepancy with lindane, which is more lipophilic than 2,4-D and least likely to be absorbed via a polar route. Despite a threefold range in follicle area in the marmo­ set, no differences in absorption rates of paraquat, mannitol, water, and ethanol were observed between different body sites (Scott et al., 1991). However, among the different spe­ cies examined in this study, there was an 80-times range in follicle area, which correlated with observed differences in the rates of mannitol and paraquat absorption. The authors concluded that this correlation was only possible with rela­ tively slowly absorbed test penetrants such as paraquat and mannitol. Further work is needed to determine the extent to which the unique anatomical features at different body sites play a role in absorption and penetration of both lipophilic and hydrophilic pesticides.

37.3.1.2  Pesticide Formulation and Mixtures Insecticide efficacy, the stability of active ingredients, and programmed release of active ingredients from the vehicle/ device are the most important characteristics controlled for when pesticides are formulated (Krenek and Rohde, 1988). EPA registration does not always require percutaneous absorption studies. For this reason, more efficacy data than dermal pharmacokinetic data are available in the litera­ ture. Furthermore, most of the available pesticide absorp­ tion data pertain to binary mixtures (pesticide  vehicle). Technical grade formulations are, however, complex mix­ tures of formulation additives and, therefore, risk assess­ ment based on data from exposure to binary mixtures may be inappropriate. Pesticides are usually formulated to contain active and inactive or inert ingredients. The latter component(s) can enhance the rate and extent of absorp­ tion or slow the release of the active ingredient and thus reduce the rate and extent of absorption (Walters and Roberts, 1993). These “inert” ingredients are often classi­ fied as adjuvants, surfactants, preservatives, solvents, dilu­ ents, thickeners, and stabilizers. These pesticide additives were first covered by the Food and Drug Administration and now are covered by EPA regulation 40 CFR 180.1001 and also TSCA and FIFRA (Seaman, 1990). This increas­ ing list of inerts as well as the prohibitive cost to obtain 40 CRF 180.1001 clearance of new inerts strongly support the

Hayes’ Handbook of Pesticide Toxicology

need to evaluate the influence of current and novel inerts on the toxicology and dermal absorption of active ingredi­ ents in pesticide formulations. Several studies have demonstrated the penetration enhancing ability of acetone compared to water, ethanol, or other vehicles commonly used in dermal absorption stud­ ies. Early work by O’Brien and Dannelley (1965) showed that in comparison with benzene and corn oil, acetone was best at enhancing carbaryl absorption. More recent stud­ ies also have demonstrated the enhancing effect of acetone compared with other solvent systems on the absorption of carbaryl, p-nitrophenol, and 2,4-D (Baynes and Riviere, 1998; Brooks and Riviere, 1995; Moody et al., 1992). However, other studies have demonstrated that com­ mercial formulations are more effective than acetone in enhancing pesticide absorption. Methyl parathion absorp­ tion in vitro in human skin at 24 h was 1.3% in acetone, but was significantly increased to 5.2% in a commercial formulation (Sartorelli et al., 1997). Likewise, in vivo dermal exposure studies of lindane in humans resulted in approximately 60% with a white spirit formulation and 5% with an acetone vehicle (Dick et al., 1997a,b). In these lat­ ter experiments, more of the lindane dose (79%) remained on the skin surface at 6 h with acetone than with the white spirit formulation (10.5%), and significant levels of lin­ dane accumulated in the stratum corneum with white spirit (30%) and with acetone (14.3%) at 6 h. These findings strongly suggest that the white spirit formulation enhanced lindane penetration with respect to acetone vehicle. The in vitro studies with human skin also demonstrated a simi­ lar pattern, although only 18 and 0.3% of the dose was absorbed into the perfusate at 6 h for the white spirit for­ mulation and the acetone vehicle, respectively. Topical application of 1% commercial lotion of lindane in vitro in human and guinea pig skin resulted in absorption levels as high as 71.72 and 35.31%, respectively, at 48-h exposure (Franz et al., 1996). Dermal absorption of alachlor as an emulsifiable con­ centrate and microencapsulated formulation was demon­ strated to be 8.5 and 3.8%, respectively, in rhesus monkeys after a 12-h exposure (Kronenberg et al., 1988). About 88% of the systemically absorbed dose was excreted in urine within 48 h. However, the differences between these two formulations were not statistically significant. Although dilution of either of these formulations (1:29) slightly enhanced alachlor absorption, these effects were surprisingly not statistically significant. One in vitro study with human skin demonstrated similar absorption data (0.5–4%) after an 8-h exposure and peak fluxes within 3–5 h postapplication (Bucks et al., 1989b). However, a significant effect of formulation dilution with water was observed in this study, even though the same mass of ala­ chlor was applied to skin. Not surprisingly, a greater frac­ tion of alachlor was present on the skin surface and skin tissue than in the receptor fluid, and the high capacity for

Chapter | 37  Absorption

stratum corneum binding demonstrated in this study is not unique for related chlorinated aromatic chemicals. Data from several studies have demonstrated that pesti­ cide applicators may be at risk to increased dermal absorption of some pesticides if they apply sunscreen or an insect repel­ lent while working with pesticides. The active ingredients in many commercial sunscreens contain UV absorbers (e.g., titanium dioxide and zinc oxide), which could act as pen­ etration enhancers of 2,4-D, paraquat, parathion, malathion (Brand et al., 2002, 2003, 2007). Other studies have demon­ strated that the active ingredient in many insect repellents, DEET, enhances transdermal delivery of drugs and toxicants (Moody et al., 1987; Windheuser et al., 1982). Some studies have demonstrated that DEET can act as a transdermal accel­ erant of 2,4-D-amine (Moody et al., 1992). Recent studies in our laboratory have, however, determined that DEET blocked permethrin absorption and inhibited carbaryl absorption in acetone, but not in dimethyl sulfoxide (DMSO) mixtures (Baynes and Riviere, 1998; Baynes et al., 1997). The insecti­ cide synergist, piperonyl butoxide, which is often formulated with some insecticidal products, was also shown to enhance carbaryl absorption (Baynes and Riviere, 1998). These dif­ fusion studies further demonstrated that piperonyl butoxide does not enhance absorption through inert latex membranes, but does so in porcine skin sections. This observation sug­ gests that some chemical–biological interaction or other mechanisms (e.g., irritation) may occur in skin to enhance the absorption of pesticides. An expected, but important find­ ing in these carbaryl experiments was that increased dilution of the carbaryl formulation with water, especially in the pres­ ence of the surfactant, sodium lauryl sulfate (SLS), enhanced carbaryl absorption. The penetration enhancing effect of SLS was also observed with parathion (Qiao et al., 1996). In addition to the formulation additives, agrochemicals may contain isomers, homologues, or breakdown products that form after synthesis and/or formulation and during storage (Chambers and Dorough, 1994). Although these impurities can potentially alter the toxicity and toxico­ kinetics of the pesticide, many toxicology and dermal absorption studies have ignored these impurities and used the pure rather than the technical grade pesticide. There is evidence that technical grade malathion can be more lethal (eightfold difference) in rats than the purified form (Umetsu et al., 1977). Other studies have demonstrated that organophosphates such as malathion and fenitrothion can potentiate the toxicity of the carbamate insecticide carbaryl (Takahashi et al., 1987). Previous metabolism studies in isolated perfused porcine skin flaps (IPPSF) (Carver et al., 1990) demonstrated a sig­ nificant first-pass metabolism of parathion to p-nitrophenol and paraoxon, and that these metabolites may be present simultaneously during absorption of parathion. Environmental exposure to parathion is never to pure parathion because spon­ taneous degradation occurs during storage. When mixtures of parathion and its metabolites were dosed and then assayed for

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parathion and its two metabolites across pig skin in vitro, sig­ nificant interactions were detected. In general, the nontoxic metabolites p-nitrophenol and 1-naphthol can significantly enhance the absorption of the parent compounds, parathion and carbaryl, respectively (Baynes and Riviere, 1998; Chang et al., 1994). Surprisingly, p-nitrophenol did not enhance the absorption of paraoxon; this toxic metabolite of parathion appears to decrease the absorption of p-nitrophenol and para­ oxon. In other related absorption studies, pretreatment with 3% fenvalerate decreased subsequent absorption of parathion, increased subsequent lindane absorption, and had no effect on subsequent fenvalerate or carbaryl absorption (Chang et al., 1995). These results underscore the chemical specificity of these interactions and reinforce the concept that the percuta­ neous absorption of a mixture cannot be predicted from indi­ vidual component studies. These data suggest that other mechanisms in addition to vehicle and surfactant effects must be operating simul­ taneously; hence further investigation is required. The data reinforce the concept that the permeability of a mixture can­ not be predicted from individual component studies. Many of the mechanisms of pesticide mixture interactions are not well understood and are not easy to model, although a bio­ physical model for parathion was attempted (Williams et al., 1996). It should also be recognized that it is more often the formulation additives and other environmental factors rather than the active ingredient that compromise the skin barrier and eventually enhance pesticide absorption. There is epi­ demiological evidence that agricultural pesticides can cause dermatoses (Abrams et al., 1991; Cellini and Offidani, 1994; Guo et al., 1996) and there is experimental evidence that UV irradiation can enhance skin reactions to topical agricul­ tural chemical treatment (Kimura et al., 1998). In the latter study, significant reactions were observed for several her­ bicides. Maibach and Feldmann (1974) demonstrated that dermal absorption of pesticides such as parathion, azodrin, and diquat occurs more readily (ninefold) through damaged skin than through normal skin. It is, therefore, plausible to assume that the formulation additive can inflict local revers­ ible or irreversible damage to the skin structure and physi­ ology, and that it is these interactions that modulate dermal absorption of most pesticides. Recent in vivo animal studies have demonstrated that oral consumption of alcohol can significantly increase the dermal absorption of the herbicides 2,4-D, paraquat, and atrazine and the insect repellent DEET by as much as 1.6- to 2.3-fold (Brand et al., 2004, 2007). The authors of these studies pro­ posed that alcohol solvates the polar head regions of the lipid in the stratum corneum and thus disrupts the interactions between the polar head group and alkyl chains. Further work by these investigators demonstrated that oral ingestion of alcohol sig­ nificantly enhanced skin lipid peroxidation and transepidermal water loss (TEWL) and this is a more plausible explanation for the increased dermal absorption of these pesticides (Brand and Jendrzejewski, 2008).

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37.3.1.3  Environmental Factors (a) Temperature Changes in ambient air temperature can alter lipid fluidity in the intercellular lipid domain of the stratum corneum. This alteration in the intercellular pathway can theoretically alter pesticide penetration through the stratum corneum. Previous in vivo studies have demonstrated that increased percutaneous absorption of a cholinesterase inhibitor (VX) was a function of skin temperature (Craig et al., 1977). In humans topically exposed to parathion at different ambi­ ent temperatures (11, 25, and 40°C), the urinary excretion of the metabolite p-nitrophenol paralleled the increase in ambient temperature (Hayes et al., 1964). Several in vitro experiments with pig skin also demonstrated that increas­ ing air temperature from 37 to 42°C significantly increased parathion absorption (Chang and Riviere, 1991). Increased ambient temperatures can also increase the evaporation of volatile pesticides from the skin, thereby reducing the topical dose available for absorption. Increasing air flow over the skin increases evaporative loss and signifi­ cantly decreases dermal residues in the upper skin layer of pigs for DDT, malathion, parathion, and DEET (Reifenrath et al., 1991). Wester et al. (1992a) demonstrated that iso­ phenfos concentrations on the human skin surface in vivo was less than 1% dose at 24 h and that evaporation from the skin surface during absorption reduced the dose available for penetration and absorption. Finally, it should be recognized that skin surface conditions in vitro are more easily con­ trolled than in vivo, and data from in vitro studies can sig­ nificantly underestimate evaporation in vivo. (b) Humidity and Occlusion Skin hydration can be increased by occlusion, with high rel­ ative humidity or immersion conditions (e.g., swimming or bathing). Although previously it was assumed that hydration changes only affect dermal absorption of polar compounds, there is significant data that suggest that at high relative humidity, this hydration effect becomes more important for nonpolar molecules such as pesticides and is most likely secondary to an increase in diffusivity of the penetrating molecule (Behl et al., 1980). Under relative humidity con­ ditions greater than 80%, parathion absorption was signifi­ cantly increased in pig skin in vitro by as much as two to three times the value under standard conditions of 60% rela­ tive humidity (Chang and Riviere, 1991, 1993). The practical application of occlusion is when pesti­ cides get into and under the clothing of workers, this cre­ ates the ideal reservoir for penetration and absorption into the skin. Occlusion can change dermal absorption by vari­ ous mechanisms, such as reducing loss of evaporation from the skin surface, enhancing skin hydration, changes in cuta­ neous metabolism, dermal irritation, and altered cutaneous blood circulation (e.g., vasodilation). Occlusion can increase hydration of the stratum corneum from as little as 5–15% to

Hayes’ Handbook of Pesticide Toxicology

as much as 50% (Bucks et al., 1989a), thereby modulating the absorption profile for the pesticide. One in vivo study with pigs (Qiao et al., 1997) demonstrated that occlusion significantly enhanced pentachlorophenol (PCP) absorption from 29.1 to 100.72% dose and changed the shape of the absorption profile in blood and plasma. The study also sug­ gested that occlusion changed the local metabolism of PCP and as a result, the 14C partitioning between plasma and red blood cells. Occlusion was also kinetically related to modi­ fication of cutaneous biotransformation of topical parathion (Qiao and Riviere, 1995). Occlusion enhanced the cutaneous metabolism of parathion to paraoxon and to p-nitrophenol as well as the percutaneous absorption and penetration of both parathion and p-nitrophenol. Occlusion also reduced para­ thion and p-nitrophenol levels in the skin, but increased p-nitrophenol and p-nitrophenol-glucuronide in the blood. Other in vivo studies (Qiao et al., 1993) showed that dermal occlusion significantly enhanced the rate and extent of parathion absorption in pigs in the abdomen (43.94 vs. 7.47%), buttocks (48.47 vs. 15.60%), back (48.82 vs. 25.00%), and shoulder (29.28 vs. 17.41%). Although sig­ nificant anatomical site differences were observed with nonoccluded skin, these site differences were concealed with occluded skin. In vitro studies with parathion also demonstrated that occlusion increased absorption from 0.46–7.69 to 1.04–17.46% at doses ranging from 4 to 400 g/cm2 (Chang and Riviere, 1993). Pesticides can be transferred from cotton fabric into and through human skin as demonstrated in several studies (Snodgrass, 1992; Wester et al., 1996b), but it should be rec­ ognized that these studies were under occlusive conditions. Dermal absorption of malathion was 3.92% with ethanol wet fabric and 0.6% with 2-day-treated cotton sheets (Wester et al., 1996b). However, malathion absorption was increased to 7.34% when the 2-day-treated/dried cotton fabric was wetted with aqueous ethanol. In the same study, absorption of glypho­ sate was 1.42% in water solution, 0.74% when applied as wet cotton sheets, and 0.08% when applied as 2-day-treated/dried cotton sheets. Absorption increased to 0.36% when the 2-daytreated/dried cotton sheets were wetted with water to simulate sweating and wet conditions. Military uniforms are impreg­ nated with permethrin as a defense against nuisance and dis­ ease-bearing insects. Application of fabric impregnated with permethrin to the backs of rabbits resulted in a 3.2% migra­ tion to the skin surface with 2% of the impregnant being absorbed and 1.2% remaining on the skin surface after 7 days of continuous skin contact (Snodgrass, 1992). The implica­ tions of these interactions, especially for agricultural workers during pesticide application in humid climates or for military personnel under combat conditions in the desert, should not be underestimated. (c) Soil Pesticide adsorption to soil can alter the amount of pesticide available for dermal absorption. It should also

Chapter | 37  Absorption

be recognized that exposure conditions such as exposure time, pesticide concentration, soil load, and soil character­ istics are important variables that can theoretically influ­ ence absorption (Bunge and Parks, 1997). Soil adherence to skin, for instance, can vary from 103 to 102 mg/cm2 and has been shown to be activity-dependent (Kissel et al., 1996). Predicting dermal absorption of pesticides from contaminated soils is, therefore, not a simple process and becomes problematic because there are very few stud­ ies that have addressed many of these issues. For several pesticides (e.g., PCP, 2,4-D, chlordane), percutaneous absorption in acetone vehicle appears to be slightly less or not significantly different from absorption from soil. However, for several other pesticides (e.g., DDT, organic arsenicals), soil appears to reduce percutaneous absorption of the pesticide. The interactions between soil and several of these pesticides are subsequently described in more detail, but note that in vitro skin models are, in general, not very predictive of in vivo absorption when exploring these interactions (Wester and Maibach, 1998). Although DDT is no longer widely used in the United States, residues in soil are still detectable and human con­ tact with contaminated soil can result in DDT exposure. One study demonstrated that in vivo absorption of DDT in rhesus monkeys was significantly less from soil (3.3% dose) than from acetone vehicle (18.9%) (Wester et al., 1990). The absorption of DDT in acetone in rhesus mon­ key is not significantly different from DDT absorption in humans (10.4% dose) (Feldmann and Maibach, 1974). In vivo absorption from acetone or soil was not similar to in vitro absorption (0.1%). However, in vitro experiments demonstrated that 18.1% penetrated skin with acetone and 1.0% penetrated skin with soil. Less than 1.0% dose par­ titioned into the receptor phase, demonstrating that the skin barrier in addition to the soil is rate-limiting and that in vitro skin models may not be useful for predicting DDT absorption in vivo. Unfortunately, only in vitro dermal absorption studies are available for organic arsenicals. One study demonstrated that as much as 12.4% dose of monosodium methyl arsenate (MSMA) and disodium methyl arsenate (DSMA) penetrated mice skin within 24 h from aqueous vehicles over a wide dosage range (Rahman and Hughes, 1994). Of this amount, only 4% were absorbed into the receptor fluid. In the pres­ ence of soil (690 ppm), penetration through mice skin was reduced to not more than 0.48 and 0.22% for MSMA and DSMA, respectively. Increasing MSMA and DSMA levels in soil from 690 to 6900 ppm increased skin content, but decreased the percentage of applied dose in skin. Whereas absorption into receptor fluid was very low for MSMA (0.01%), it was not detectable for DSMA. Topical applica­ tion of aqueous solutions (20, 100, and 250 l) of 10 g of DMA to mice skin resulted in 5.16–25.22% dose in recep­ tor fluid and 1.95–15.67% dose in skin tissue within 24 h (Hughes et al., 1995). However, when DMA (690 ppm) was

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applied with soil, absorption was reduced to 0.08% in the receptor fluid and 0.45% in skin. The influence of soil was, however, not observed with inorganic arsenic. In vivo percutaneous absorption of arse­ nic as H3A5O4 in water in rhesus monkeys (2.0–6.4%) was somewhat comparable to in vitro absorption (1.9%) in human skin (Wester et al., 1993a). However, the soil vehicle did not influence absorption in rhesus monkeys (3.2–4.5%) or human skin in vitro (0.8%), although absorption in both skin models is not comparable. The relative similarities in the partition coefficient of arsenic from water to stra­ tum corneum and from water to soil probably explain why absorption from water was similar to absorption from soil. Interactions between soil and the phenoxy herbicides (e.g., 2,4-D acid, 2,4-D amine) are unique. One study dem­ onstrated that dermal absorption of the herbicide 2,4-D acid is nonlinear with respect to soil load or skin contact time (Wester et al., 1996a). Percutaneous absorption in acetone vehicle (8.6%) was not different from absorption of soil loads of 1 mg/cm2 (8.6%) and 40 mg/cm2 (15.9%) in rhesus monkey in vivo. Further in vitro experiments with human skin demonstrated that increasing the soil load from 5 to 40 mg/cm2 did not affect 2,4-D absorption, which ranged from 1.4 to 1.8%. During the first 24 h of in vitro exposure, absorption was linear with respect to time for an acetone vehicle (3.2%); however, there was an apparent lag time of about 8 h with absorption from a soil vehicle (0.03–0.05%). This early lag time may be related to chemi­ cal partitioning from soil and may be beneficial if the skin is decontaminated within 24 h. The investigators proposed that because of complex interactive forces between pesti­ cides and soil, dermal absorption calculations based on assumed linearity can incorrectly estimate the threat to human health. Mathematical extrapolation from high soil loads to low soil loads may significantly underestimate 2,4-D absorption. These studies also demonstrated that soil release kinetics may limit dermal absorption and that more data are needed to make valid predictions. It is therefore plausible to assume that only pesticides in the soil layer that is in direct contact with skin is bioavailable and heavy soil loads may not necessarily increase dermal absorption. In contrast to DDT, chlordane absorption in rhesus monkeys in acetone (6.0% dose) was similar to absorption in soil (4.2% dose) 6 days after exposure (Wester et al., 1992b). Although human skin in vitro experiments dem­ onstrated similar partitioning into receptor fluid for ace­ tone (0.07%) and soil vehicles (0.04%), there was greater penetration into skin with acetone (10.8%) than with soil (0.34% dose) at 24 h. It is possible that chlordane adsorp­ tion to soil delayed percutaneous absorption during the initial 24 h and an extrapolation to 6 days would reveal no vehicle differences as demonstrated in the in vivo study. The octanol:water partitioning coefficients (log P) of chlordane and DDT are 5.58 and 6.91, respectively, and, therefore, dermal disposition should be similar. The high

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lipophilicity of these pesticides explains the higher propor­ tion of pesticide in the skin than in the receptor phase, but it does not explain why the differences between acetone and soil for DDT are greater than those for chlordane; it only suggests that factors other than lipophilicity influence absorption of these organochlorines. Various studies have demonstrated that the very ubiq­ uitous pesticide, PCP, is very readily absorbed through human, monkey, and pig skin (Qiao et al., 1997; Wester et al., 1993b). In vivo absorption of PCP in rhesus monkeys with acetone vehicle (29.2%) was similar to absorption with soil vehicle (24.4%) after a 24-h exposure period (Wester et al., 1993b). However, in vitro absorption with human skin appears to underestimate in vivo absorption because only 0.6–1.5 and 0.01–0.07% dose were detected in recep­ tor fluid with acetone and soil vehicle, respectively, at 24 h. Skin concentrations were only 2.6–3.7 and 0.11–0.14% for acetone and soil vehicles, respectively, which is still not comparable to the in vivo data. The approximately 25% PCP absorption from nonocclusive soil in monkeys (Wester et al., 1993b) compares favorably with the 29% PCP absorption from nonocclusive soil in Yorkshire pigs in vivo (Qiao et al., 1997). Note that inhibition of soil and/or skin microorganisms can inhibit absorption of PCP, alter local and systemic dis­ tribution, and increase plasma/blood concentration ratios in pig skin in vivo (Qiao et al., 1997). It is plausible to assume that skin or soil microorganisms and/or products of PCP microbial degradation may play a role in PCP absorp­ tion and disposition.

37.3.2  Absorption From the   Gastrointestinal Tract People are potentially exposed to pesticides by oral expo­ sure from pesticide residues in foods such as meat, milk, fruits, and vegetables. Children may also be orally exposed to pesticides when they place contaminated objects in their mouths. The rate and extent of absorption after oral exposure depends on the ability of the chemical to cross the plasma membranes of the gastrointestinal tract. As discussed in Section 37.2, diffusion across a plasma mem­ brane depends to a large extent on the lipid solubility and degree of ionization of the chemical. The degree of ioniza­ tion of weak acids and weak bases, and hence absorption, depends on pH. The pH range in the gastrointestinal tract varies from approximately 1–3 in the stomach to 6–8 in the intestines. Thus, the rate and extent of absorption of weak organic acids and bases varies with location in the gastro­ intestinal tract; weak acids are nonionized and are absorbed in the stomach, whereas weak bases are nonionized and are absorbed in the intestine (Figure 37.4). Removal from the site of absorption by blood flow maintains a concentration gradient, thus enhancing absorption of chemicals.

Hayes’ Handbook of Pesticide Toxicology

Residence time in the region of the gastrointestinal tract where the chemical is absorbed also affects absorption. The presence of food in the gut can alter the pH of the gut con­ tents and the intestinal motility, which in turn can affect the rate of absorption from the gastrointestinal tract. The pres­ ence of stomach acid, gastric enzymes, and intestinal flora may decompose the chemical before absorption can occur, which may also decrease the potential for toxicity. Pekas (1972) demonstrated the hydrolysis of naphthyl N-methyl carbamate in the intestine (pH 6.4). The large surface area of the intestinal tract aids in absorption from this site; even chemicals that do not readily cross the plasma membrane (e.g., weak acids) can be absorbed to a high degree in the intestine because of the increased surface area. Particles may be absorbed in the intestines by endocytosis. Chemicals that are absorbed into the bloodstream from the gastrointestinal tract enter the portal circulation and are delivered directly to the liver, where they may be metabolized before reaching the systemic circulation. This first-pass metabolism decreases the systemic avail­ ability (bioavailability) of the parent compound. Some highly lipophilic compounds such as organochlorine pesticides are absorbed into the lymphatic system in a manner similar to the absorption of nutritional fats (Turner and Shanks, 1980). Absorption into the lymphatic sys­ tem bypasses delivery to the liver and the potential for first-pass metabolism. Many of the in vivo oral absorption studies in laboratory animals have demonstrated that oral absorption of many pesticides can be significantly greater than absorption fol­ lowing skin exposure. This is especially applicable to the pyrethroid class of insecticides, which do not readily pen­ etrate skin and usually have a dermal bioavailability of less than 10% in human skin. Deltamethrin is rapidly absorbed (Tmax 5 1.0 h) and its absolute oral bioavailability was 18% in rats given doses in the range of 0.2–10 mg/kg by oral gavage in glycerol formal (Kim et al., 2007). Per­ methrin was less rapidly absorbed (Tmax  3.52 h) and had a greater absolute bioavailability (F  61%) in the same strain of rats (Anadon et al., 1991). Many of the organophosphates are readily absorbed, although the data suggest that pesticide formulation can result in absorption varying from 20% to as much as 70% according to sev­ eral human studies (Eaton et al., 2008; Nolan et al., 1984; Timchalk et al., 2002). The commonly used chlorophenoxy herbicide, 2,4-D, is rapidly absorbed by the human gastro­ intestinal tract (Kohli et al., 1974; Sauerhoff et al., 1977); however, a more recent rodent study suggests that clear­ ance can be uniquely sex-dependent (Griffin et al., 1997). The quarternary nitrogen herbicide, paraquat, has a limited bioavailability in laboratory animals of about 22% and less than 5% absorption during the first 6 h (Chui et al., 1988; Meredith and Vale, 1987), although it has often been associated with human poisonings following accidental or suicidal ingestion. Paraquat absorption across

Chapter | 37  Absorption

887

Stomach contents (pH 1.0) Nonionized/ionized ratio favors absorption Blood (pH 7.4)

Ionized [54,000]

Nonionized [1]

Ionized [0.43]

Nonionized [1]

Ionized [460]

Nonionized [1]

Intestinal contents (pH 5.3) Nonionized/ionized ratio does not favor absorption

the GIT can be rapid, with the peak plasma concentration occurring within the first hour after exposure (Heylings et al., 1991; Nagao et al., 1993) and it is plausible to assume that the presence of herbicide performance surfac­ tants may play a significant role in paraquat uptake across the GIT. There have been recent successful efforts to include additives such as alginate to paraquat formulations to reduce paraquat uptake by the GIT of experimental ani­ mals. The mechanism involves delaying gastric emptying, which prevents early high lethal dose delivery to the lungs and thus enhances human survival following acute expo­ sure to this lethal herbicide (Heylings et al., 2007). Atrazine, which belongs to the class of triazine herbi­ cides, displays unique absorption kinetics that is not often observed with many pesticides. While one rodent study has estimated that oral absorption may be slow with an absorp­ tion half-life of 3 h and limited to approximately 57% based on 24-h urine excretion data (Timchalk et al., 1990), more recent rodent studies demonstrated a double peak phenom­ enon as evidenced by a single first-order absorption process and a longer plateau (McMullin et al., 2003, 2007). These studies also estimated that limited solubility in the intestine and presystemically metabolized atrazine in the intestine may account for limited oral bioavailability. However, it should be emphasized again that oral absorption is signifi­ cantly greater by oral exposure than by dermal exposure as exemplified here by atrazine, where only as much as 5.6% of the topical dose is expected to be absorbed across the skin of humans and as much as ten times this amount could be absorbed by the gastrointestinal tract. Risk assessors should be wary of formulation and dose effects, which will alter bioavailability across both routes of exposure.

Figure 37.4  Proportion of nonionized and ionized forms of 2,4-D (pKa  2.6) in the stomach and intestinal contents (adapted from Hodgson et al., 1991). Only the non­ ionized form crosses cell membranes.

37.3.3  Absorption From the   Respiratory Tract A chemical must be in the form of a gas, vapor, or particulate (e.g., aerosol) to be absorbed in the respiratory tract. Although the anatomy of the respiratory tract varies widely within mammalian species, the respiratory system can be generally compartmentalized into the nasopharyngeal, tracheobron­ chial, and pulmonary regions (Kennedy and Valentine, 1994). The function of the nasopharyngeal region is to condition inspired air and to remove large inspired particles before they reach the tracheobronchial and pulmonary regions. The tra­ cheobronchial region is lined by mucus-se­creting and ciliated cells, which together make up the mucociliary escalator. The pulmonary region of the respiratory tract is the gas-exchange region, which consists of the respiratory bronchioles, alveolar ducts, and alveoli. Inspired gases and vapors may be absorbed throughout the respiratory tract, depending on their physicochemical properties, and the anatomy and physiology of the region. Inhaled gases and vapors diffuse across cell membranes in the direction of the concentration gradient until equilibrium is established (see also Section 37.2.2). The ratio of gas or vapor equilibrium concentrations in blood and air is termed the blood:air partition coefficient. Highly water-soluble and reactive gases and vapors tend to be absorbed in the mucus layer of the upper respiratory tract, whereas more lipo­ philic and nonreactive gases and vapors are absorbed from the deeper regions of the respiratory tract. The geometry, blood flow, and capacity for metabolism of the respiratory tract may also influence the rate and site of absorption of inhaled gases and vapors.

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Deposition of aerosols in the respiratory tract depends on a number of factors, including the physicochemical and aerodynamic properties of the aerosol and the geometry and airflow of the respiratory tract. Airflow (velocity) and turbulence decrease from the nasopharyngeal to the pulmo­ nary region, and different mechanisms of deposition oper­ ate in these regions. Impaction is an important mechanism for deposition of particles larger than 1 m in aerodynamic diameter in regions of the respiratory tract where air velo­ city and turbulence are high, such as airway bifurcations. Interception is an important mechanism of deposition for fibers and is dependent on fiber length rather than dia­ meter. Sedimentation is settling due to gravity and is impor­ tant for particles larger than 1 m in aerodynamic diameter in regions of the respiratory tract where airways are small in diameter and airflow is low. Diffusion is an important mechanism of deposition for small particles throughout the respiratory tract, and particularly for particles 0.5 m in the alveolar region where airflow is low. Particles are cleared from the respiratory tract in a num­ ber of ways, depending on the region of the respiratory tract. In the nasopharyngeal region, particles are removed by nose wiping, nose blowing, sneezing, and swallowing. In the tra­ cheobronchial region, particles are cleared by the mucocili­ ary escalator and are ultimately swallowed or expectorated. In the pulmonary region, particles are removed by (1) dis­ solution and removal in the bloodstream or lymphatics, (2) alveolar macrophage phagocytosis and removal by the muco­ ciliary escalator or lymphatics, and (3) direct penetration of epithelial membranes and absorption into tissue or blood.

37.3.4  Absorption After Exposure by   Other Routes Intravenous (IV) administration, in which the chemical is introduced directly into the blood, by definition, bypasses the process of absorption. The advantages of administer­ ing a chemical by this route include rapid achievement of effective blood concentrations, precise knowledge of the delivered dose, and the ability to deliver the chemical that would cause irritation by other routes. As will be discussed in the pharmacokinetics section of this book, the plasma concentration–time data obtained from IV administration is necessary for estimating the absolute bioavailability of a substance given by an extravascular route. The major dis­ advantages of IV administration are that the administered dose cannot be removed, systemic toxicity may occur with some pesticides because of the transiently high concentration achieved, sterile pesticide preparations are required, and pes­ ticides that are insoluble in plasma may precipitate and cause emboli formation. For experimental pharmacokinetic studies, pesticides can be administered intravenously by bolus injec­ tion, infusion, or added to fluid drip bags. Intra-arterial (IA) injection is very similar to IV administration, but

Hayes’ Handbook of Pesticide Toxicology

more hazardous since the pesticide is delivered directly to the arterial circulation. This mode is often used in pharma­ cology to treat localized tumors that have accessible arter­ ies for injection. Other routes of exposure that are used in the laboratory include intraperitoneal (IP) injection, where chemicals are absorbed primarily through the portal circulation, and intra­ muscular or subcutaneous injection. Chemicals administered intramuscularly (IM) or subcutaneously (SC) are absorbed more slowly than by IP injection. Injection is into the mus­ cle mass, which is well perfused by the vascular system, and can result in usually rapid absorption. Lipid solubility of the administered drug is not important as both hydrophilic and charged chemicals are easily absorbed in the rich capillary networks of muscle tissue. Absorption may be modulated by the vehicle used to inject the drug into the muscle. In this case, release of drug from the vehicle becomes the ratelimiting step. The SC mode of administration is very simi­ lar to IM injection except that the drug is deposited into the rich capillary beds perfusing the skin and absorption is (in general) slower than the IM route and can be very erratic. Injection of tissue-irritating substances may cause local reac­ tions and even skin sloughing, which alter absorption. Other routes of administration are intradermal, intramammary, and subconjunctival, which are not often associated with animal testing or human routes of exposure for pesticides.

37.4  Summary and Future Directions Human exposure to pesticides can result in absorption by passive diffusion across the epithelium layers of the skin and/or the gastrointestinal tract. As occupational exposure via the skin is the most likely source and route for human exposure, this chapter focused on the main physicochemi­ cal, biological, and environmental factors that could sig­ nificantly influence systemic absorption following dermal exposure to examples of pesticides with diverse physico­ chemical properties. There are several in vitro and inert membrane models (Bronaugh and Stewart, 1984; Carver et al., 1989) that can be used experimentally to evaluate membrane transport and there has been some success in comparing absorption across in vitro and in vivo systems (Wester et al., 1998). These in vitro systems can be used to assess the relative influence of several formulation and bio­ logical factors that determine dermal and oral absorption of pesticides without having to use in vivo animal models or humans in the early stages of formulation development or human health risk assessments. The recent improvements in computer modeling capabilities have resulted in devel­ opment of numerous quantitative structure permeability relationships (QSPRs) that have proven to be predictive of dermal and oral permeability in humans for several solutes, including many pesticides currently in use today (Baynes

Chapter | 37  Absorption

et al., 2008; Potts and Guy, 1992; Riviere and Brooks, 2007; Zhao et al., 2002). A very popular method, the rule of 5, utilizes similar QSPR principles, and has proven to be useful as a rapid screen for compounds that are likely to be poorly absorbed orally (Lipinski et al., 1997). This rule states that if a compound satisfies any two of the following rules, it is likely to exhibit poor intestinal absorp­ tion: (1) molecular weight 500, (2) number of hydrogen bond donors 5 (a donor being any O-H or N-H group), (3) number of hydrogen acceptors 10 (an acceptor being any O or N including those in donor groups), and (4) C log P  5.0 or M log P  4.15. There are numerous other orig­ inal peer-reviewed research articles that readers are encour­ aged to review to get a glimpse of how these QSPR models are being developed and applied to human risk assessment. It should be noted that, in many instances, many of the data used in developing these models are often obtained from multiple sources and laboratories with diverse dos­ ing and experimental protocols and the statistical analyses (e.g., appropriate measures of goodness-of-fit, robustness, and predictivity) and a defined application domain may not be reported. Many of these criteria are important for valid evaluation of QSAR models that could predict pes­ ticide absorption as new regulatory frameworks such as REACH in the EU are implemented in various jurisdictions (Bouwman et al., 2008). Caution should therefore be exer­ cised with the mechanistic interpretation and application of these permeability models to any given exposure scenario. The future of risk assessment of pesticides will depend heavily on the quality of these permeability models and their flexibility to predict dermal and oral absorption in a variety of human exposure scenarios at home and work.

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Behl, C. R., Flynn, G. L., Kurihara, T., Harper, N., Smith, H., Higuchi, W. I., Ho, N. F. H., and Pierson, C. (1980). Hydration and percutaneous absorption: I. Influence of hydration on alkanol permeation through hairless mouse skin. J. Invest. Dermatol. 75, 346–352. Bouwman, T., Cronin, M. T., Bessems, J. G., and van de Sandt, J. J. (2008). Improving the applicability of (Q)SARs for percutaneous penetration in regulatory risk assessment. Hum. Exp. Toxicol. 27(4), 269–276. Brand, R. M., and Jendrzejewski, J. L. (2008). Chronic ethanol ingestion alters xenobiotic absorption through the skin: potential role of oxida­ tive stress. Food Chem. Toxicol. 46(6), 1940–1948. Brand, R. M., Spalding, M., and Mueller, C. (2002). Sunscreens can increase dermal penetration of 2,4-dichlorophenoxyacetic acid. J. Toxicol. Clin. Toxicol. 40(7), 827–832. Brand, R. M., Pike, J., Wilson, R. M., and Charron, A. R. (2003). Sunscreens containing physical UV blockers can increase transder­ mal absorption of pesticides. Toxicol. Ind. Health 19(1), 9–16. Brand, R. M., Charron, A. R., Dutton, L., Gavlik, T. L., Mueller, C., Hamel, F. G., Chakkalakal, D., and Donohue, T. M. Jr. (2004). Effects of chronic alcohol consumption on dermal penetration of pesticides in rats. J. Toxicol. Environ. Health A 67(2), 153–161. Brand, R. M., Jendrzejewski, J. L., and Charron, A. R. (2007). Potential mechanisms by which a single drink of alcohol can increase trans­ dermal absorption of topically applied chemicals. Toxicology 235(3), 141–149. Bronaugh, R. L. (1985). In vitro methods for percutaneous absorption of pesticides. In “Dermal Exposure to Pesticide Use” (R. Honeycut, G. Zweig, and N. N. Ragsdale, eds.). American Chemical Society, Washington, DC. Brooks, J. D., and Riviere, J. E. (1995). Quantitative percutaneous absorp­ tion and cutaneous distribution of binary mixtures of phenol and p-nitrophenol in isolated perfused porcine skin. Fundam. Appl. Toxicol. 32, 233–243. Brunet, J. L., Maresca, M., Fantini, J., and Belzunces, L. P. (2008). Intestinal absorption of the acetamiprid neonicotinoid by Caco-2 cells: transepithelial transport, cellular uptake and efflux. J. Environ. Sci. Health B 43(3), 261–270. Bucks, D. A. W., Maibach, H. I., and Guy, R. H. (1989a). Occlusion does not uniformly enhance penetration in vivo. In “Percutaneous Absorption” (R. L. Bronaugh and H. I. Maibach, eds.). Dekker, New York. Bucks, D. A. W., Wester, R. C., Mobayen, M. M., Yang, D., Maibach, H. I., and Coleman, D. L. (1989b). In vitro percutaneous absorption and stratum corneum binding of alachlor: Effect of formulation dilution with water. Toxicol. Appl. Pharmacol. 100, 417–423. Bunge, A. L., and Parks, J. M. (1997). Predicting dermal absorption from contact with chemically contaminated soils. In “Environmental Toxicology and Risk Assessment: Modelling and Risk Assessment” (F. J. Dwyer, T. R. Doane, and M. L. Hinman, eds.), Vol. 6. ASTM STP 1317, American Society for Testing and Materials, Philadelphia. Carver, M. P., Williams, P. L., and Riviere, J. E. (1989). The isolated per­ fused porcine skin flap. III. Percutaneous absorption pharmacokinetics of organophosphates, steroids, benzoic acid, and caffeine. Toxicol. Appl. Pharmacol. 97, 324–337. Carver, M. P., Levi, P. E., and Riviere, J. E. (1990). Parathion metabol­ ism during percutaneous absorption in perfused porcine skin. Pest. Biochem. Physiol. 38, 245–254. Cellini, A., and Offidani, A. (1994). An epidemiological study on cuta­ neous diseases of agricultural workers authorized to use pesticides. Dermatology 189, 129–132. Chambers, J. E., and Dorough, G. D. (1994). Toxicological prob­ lems associated with pesticide mixtures and pesticide impurities.

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In “Toxicology of Chemical Mixtures” (R. H. S. Yang, ed.), pp. 135–155. Academic Press, San Diego. Chang, S. K., and Riviere, J. E. (1991). Percutaneous absorption of para­ thion in vivo in porcine skin. Effects of dose, temperature, humidity and perfusate composition on absorptive flux. Fundam. Appl. Toxicol. 17, 494–504. Chang, S. K., and Riviere, J. E. (1993). Effect of humidity and occlusion on the percutaneous absorption of parathion in vitro. Pharm. Res. 10, 152–155. Chang, S. K., Dauterman, W. C., and Riviere, J. E. (1994). Percutaneous absorption of parathion and its metabolites, paraoxon and p-nitro­ phenol, administered alone or in combination: In vitro flow-through diffusion cell studies. Pestic. Biochem. Physiol. 48, 56–62. Chang, S. K., Brooks, J. D., Monteiro-Riviere, N. A., and Riviere, J. E. (1995). Enhancing or blocking effect of fenvalerate on the subse­ quent percutaneous absorption of pesticides in vitro. Pest. Biochem. Physiol. 51, 214–219. Chui, Y. C., Poon, G., and Law, F. (1988). Toxicokinetics and bioavail­ ability of paraquat in rats following different routes of administration. Toxicol. Indust. Health 4, 203–219. Cleek, R. L., and Bunge, A. L. (1993). A new method for estimating der­ mal absorption from chemical exposure. 1. General approach. Pharm. Res. 10, 497–506. Craig, F. N., Cummings, E. G., and Sim, V. M. (1977). Environmental temperature and the percutaneous absorption of cholinesterase inhibi­ tor, VX. J. Invest. Dermatol. 68, 357–361. Dick, I. P., Blain, P. G., and Williams, F. M. (1997a). The percutaneous absorption and skin distribution of lindane in man. I. In vivo studies. Hum. Exp. Toxicol 16, 645–651. Dick, I. P., Blain, P. G., and Williams, F. M. (1997b). The percutaneous absorption and skin distribution of lindane in man. II. In vitro studies. Hum. Exp. Toxicol. 16, 652–657. Eaton, D. L., Daroff, R. B., Autrup, H., Bridges, J., Buffler, P., Costa, L. G., Coyle, J., McKhann, G., Mobley, W. C., Nadel, L., Neubert, D., Schulte-Hermann, R., and Spencer, P. S. (2008). Review of the toxi­ cology of chlorpyrifos with an emphasis on human exposure and neu­ rodevelopment. Crit. Rev. Toxicol. 38, 1–125. Elias, P. M. (1981). Epidermal lipids, membranes, and keratinization. Int. J. Dermatol. 20, 1–19. Feldmann, R. J., and Maibach, H. I. (1974). Percutaneous penetration of some pesticides and herbicides in man. Toxicol. Appl. Pharmacol. 28, 126–132. Ferrer, A., and Cabral, R. (1993). Collective poisonings caused by pes­ ticides: Mechanism of production–mechanism of prevention. In “Reviews in Environmental Toxicology” (E. Hodgson, ed.), Vol. 5, pp. 161–201. Toxicology Communications Inc, Raleigh, NC. Franz, T. J., Lehman, P. A., Franz, S. F., and Guin, J. D. (1996). Comparative percutaneous absorption of lindane and permethrin. Arch. Dermatol. 132, 901–905. Galli, C. L., and Marinovich, M. (1987). Dermal toxicity of pesticides. In “Toxicology of Pesticides: Experimental, Clinical and Regulatory Perspectives” (L. G. Costa, C. L. Galli, and S. D. Murphy, eds.), pp. 147–169. Springer-Verlag, New York. Garrettson, L. K., and Curley, A. (1969). Dieldrin: Studies in a poisoned child. Arch. Environ. Health 19, 814–822. Gomez-Catalan, J., To-Figueras, J., Rodamilans, M., and Corbella, J. (1991). Transport of organochlorine residues in the rat and human blood. Arch. Environ. Contam. Toxicol. 20, 61–66. Griffin, R. J., Godfrey, V. B., Kim, Y. C., and Burka, L. T. (1997). Sexdependent differences in the disposition of 2,4-dichlorophenoxyacetic

Hayes’ Handbook of Pesticide Toxicology

acid in Sprague-Dawley rats, B6C3F1 mice, and Syrian hamsters. Drug Metab. Dispos. 25(9), 1065–1071. Guo, Y. L., Wang, B. J., Lee, C. C., and Wang, J. D. (1996). Prevalence of dermatoses and skin sensitization associated with use of pesticides in fruit farmers of southern Taiwan. Occup. Environ. Med. 53, 427–431. Hayes, W. J., Funckes, A. J., and Hartwell, W. V. (1964). Dermal expo­ sure of human volunteers to parathion. Arch. Environ. Health 8, 829–833. Heylings, J. R. (1991). Gastrointestinal absorption of paraquat in the iso­ lated mucosa of the rat. Toxicol. Appl. Pharmacol. 107(3), 482–943. Heylings, J. R., Farnworth, M. J., Swain, C. M., Clapp, M. J., and Elliott, B. M. (2007). Identification of an alginate-based formulation of para­ quat to reduce the exposure of the herbicide following oral ingestion. Toxicology 241(1-2), 1–10. Hodgson, E., Silver, I. S., Butler, L. E., Lawton, M. P., and Levi, P. E. (1991). Metabolism. In “Handbook of Pesticide Toxicology” (W. J. Hayes Jr. and E. R. Laws Jr., eds.), pp. 107–167. Academic Press, San Diego. Hughes, M. F., Mitchel, C. T., Edwards, B. C., and Rahman, M. S. (1995). In vitro percutaneous absorption of dimethylarsinic acid in mice. J. Toxicol. Environ. Health 45, 279–290. Kennedy, G. L. Jr., and Valentine, R. (1994). Inhalation toxicology. In “Principles and Methods of Toxicology” (A. W. Hayes, ed.), 3rd ed., pp. 805–838. Raven Press, New York. Kim, K. B., Anand, S. S., Kim, H. J., White, C. A., and Bruckner, J. V. (2007). Toxicokinetics and tissue distribution of deltamethrin in adult Sprague-Dawley rats. Toxicol. Sci. 101(2), 197–205. Kimura, T., Kuroki, K., and Doi, K. (1998). Dermatotoxicity of agricul­ tural chemicals in the dorsal skin of hairless dogs. Toxicol. Pathol. 26, 442–447. Kissel, J. C., Richter, K. Y., and Fenske, R. A. (1996). Field measurement of dermal soil loading attributable to various activities: Implications for exposure assessment. Risk Anal. 16, 115–125. Kohli, J. D., Khanna, R. N., Gupta, B. N., Dhar, M. M., Tandon, J. S., and Sircar, K. P. (1974). Absorption and excretion of 2,4-dichlorophen­ oxyacetic acid in man. Xenobiotica 4, 97–100. Krenek, M. R., and Rohde, W. H. (1988). An overview – Solvents for agricultural chemicals. In “Pesticide Formulations and Application Systems” (D. A. Hovde and D. Beestman, eds.), pp. 113–127. American Society for Testing and Materials, Ann Arbor, MI. Kronenberg, J. M., Fuhremann, T. W., and Johnson, D. E. (1988). Percutaneous absorption and excretion of alachlor in rhesus monkeys. Fundam. Appl. Toxicol. 10, 664–671. Lipinski, C. A., Lombardo, F., Dominy, B. W., and Feeney, P. J. (1997). Experimental and computational approaches to estimate solubility and permeability in drug discovery and development settings. Adv. Drug Deliv. Rev. 23, 3–25. Magee, P. (1991). Percutaneous absorption: critical factors in transdermal transport. In “Dermatoxicology” (F. N. Marzulli and H. I. Maibach, eds.), pp. 1–36. Hemisphere Publishing, New York. Maibach, H., and Feldmann, R. (1974). Occupational Exposure to Pesticides. Report to the Federal Working Group on Pest Management for the Task Group, pp. 120–127. Maibach, H. I., Felman, R. J., Milby, T. H., and Serat, W. F. (1971). Regional variation in percutaneous penetration in man. Arch. Environ. Health 23, 208–211. Martin-Bouyer, G., Khanh, N. B., Linh, P. D., Hoa, D. Q., Tuan, L. C., Tourneau, J., Barin, C., Guerbois, H., and Binh, T. V. (1983). Epidemic of haemorrhagic disease in Vietnamese infants caused by warfarin contaminated talcs. Lancet 1, 230–232.

Chapter | 37  Absorption

McMullin, T. S., Brzezicki, J. M., Cranmer, B. K., Tessari, J. D., and Andersen, M. E. (2003). Pharmacokinetic modeling of disposition and time-course studies with [14C]atrazine. J. Toxicol. Environ. Health A 66(10), 941–964. McMullin, T. S., Hanneman, W. H., Cranmer, B. K., Tessari, J. D., and Andersen, M. E. (2007). Oral absorption and oxidative metabolism of atrazine in rats evaluated by physiological modeling approaches. Toxicology 240, 1–14. Meredith, T. J., and Vale, J. A. (1987). Treatment of paraquat poisoning in man: methods to prevent absorption. Hum. Toxicol. 6(1), 49–55. Monteiro-Riviere, N. A., Bristol, D. G., Manning, T. O., and Riviere, J. E. (1990). Interspecies and interregional analysis of the comparative histological thickness and laser Doppler blood flow measurements at five cutaneous sites in nine species. J. Invest. Dermatol. 95, 582–586. Moody, R. P., and Franklin, C. A. (1987). Percutaneous absorption of the insecticides fenitrothion and aminocarb in rats and monkeys. J. Toxicol. Environ. Health 20, 209–218. Moody, R. P., and Ritter, L. (1989). Dermal absorption of the insecticide lindane (1a,2a,3b,4a,5a,6b-hexachlorocyclohexane) in rats and rhe­ sus monkeys: Effect of anatomical site. J. Toxicol. Environ. Health 28, 161–169. Moody, R. P., Riedel, D., Ritter, L., and Franklin, C. A. (1987). The effect of DEET (N,N-diethyl-m-toluamide) on dermal persistence and absorption of the insecticide fenitrothion in rats and monkeys. J. Toxicol. Environ. Health 22, 471–479. Moody, R. P., Benoit, F. M., Riedel, D., and Ritter, L. (1989). Dermal absorption of the insect repellent DEET (N,N-diethyl-m-toluamide) in rats and monkeys: Effect of anatomical site and multiple exposure. J. Toxicol. Environ. Health 26, 137–147. Moody, R. P., Franklin, C. A., Ritter, L., and Maibach, H. I. (1990). Dermal absorption of the phenoxy herbicides 2,4-D, 2,4-D amine, 2,4-D isooc­ tyl, and 2,4,5-T in rabbits, rats, rhesus monkeys, and humans: A crossspecies comparison. J. Toxicol. Environ. Health 29, 237–245. Moody, R. P., Wester, R. C., Melendres, J. L., and Maibach, H. I. (1992). Dermal absorption of the phenoxy herbicde 2,4-D dimethylamine in humans: Effect of DEET and anatomic site. J. Toxicol. Emviron. Health 36, 241–250. Nagao, M., Saitoh, H., Zhang, W. D., Iseki, K., Yamada, Y., Takatori, T., and Miyazaki, K. (1993). Transport characteristics of paraquat across rat intestinal brush-border membrane. Arch. Toxicol. 67(4), 262–267. Nolan, R. J., Rick, D. L., Freshour, N. L., and Saunders, J. H. (1984). Chlorpyrifos: Pharmacokinetics in human volunteers. Toxicol. Appl. Pharmacol. 73, 8–15. O’Brien, R. D., and Dannelley, C. E. (1965). Penetration of insecticides through rat skin. J. Agric. Food Chem. 13, 245–247. Opdycke, J. C., and Menzer, R. E. (1984). Pharmacokinetics of diflubenzuron in two types of chickens. J. Toxicol. Environ. Health 13, 721–733. Pekas, J. C. (1972). Intestinal hydrolysis, metabolism and transport of a pesticidal carbamate in pH 6.5 medium. Toxicol. Appl. Pharmacol. 23, 62–70. Potts, R. O., and Guy, R. H. (1992). Predicting skin permeability. Pharm. Res. 9, 663–669. Qiao, G. L., and Riviere, J. E. (1995). Significant effects of application site and occlusion on the pharmacokinetics of cutaneous penetration and biotrans-formation of parathion in vivo in swine. J. Pharm. Sci. 84, 425–432. Qiao, G. L., Chang, S. K., and Riviere, J. E. (1993). Effects of anatomical site and occlusion on the percutaneous absorption and residue pattern of 2,6-[ring-14C]parathion in vivo in pigs. Toxicol. Appl. Pharmacol. 122, 131–138.

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Qiao, G. L., Brooks, J. D., Baynes, R. E., Monteiro-Riviere, N. A., Williams, P. L., and Riviere, J. E. (1996). The use of mechanistically defined chemical mixtures (MDCM) to assess component effects on the percutaneous absorption and cutaneous disposition of topically exposed chemicals. Toxicol. Appl. Pharmacol. 141, 473–486. Qiao, G. L., Brooks, J. D., and Riviere, J. E. (1997). Pentachlorophenol dermal absorption and disposition from soil in swine. Effects of occlusion and skin microorganism inhibition. Toxicol. Appl. Pharmacol. 147, 234–246. Rahman, M. S., and Hughes, M. F. (1994). In vitro percutaneous absorp­ tion of monosodium methanearsonate and disodium methanearsonate in female B6C3F1 mice. J. Toxicol. Environ. Health 41, 421–433. Reifenrath, W. G., Hawkins, G. S., and Kurtz, M. S. (1991). Percutaneous penetration and skin retention of topically applied compounds: An in vitro–in vivo study. J. Pharm. Sci. 80, 526–532. Riviere, J. E., and Brooks, J. D. (2007). Prediction of dermal absorption from complex chemical mixtures: incorporation of vehicle effects and interactions into a QSPR framework. SAR QSAR Environ Res. 2007 Jan-Mar 18(1-2), 31–44. Sartorelli, P., Aprea, C., Bussani, R., Novelli, M. T., Orsi, D., and Sciarra, G. (1997). In vitro percutaneous penetration of methyl-parathion from a commercial formulation through the human skin. Occupat. Environ. Med. 54, 524–525. Sauerhoff, M. W., Braun, W. H., Blau, G. E., and Gehring, P. J. (1977). The fate of 2,4-dichlorophenoxyacetic acid (2,4-D) following oral administration to man. Toxicology 8, 3–11. Scheuplein, R. J. (1972). Properties of the skin as a membrane. Adv. Biol. Skin 12, 125–152. Scott, R. C., Corrigan, M. A., Smith, F., and Mason, H. (1991). The influ­ ence of skin structure on permeability: An intersite and interspecies comparison with hydrophilic penetrants. J. Invest. Dermatol. 96, 921–925. Seaman, D. (1990). Trends in the formulation of pesticides. Pestic. Sci. 29, 437–449. Sidon, E. W., Moody, R. P., and Franklin, C. A. (1988). Percutaneous absorp­ tion of cis-and trans-permethrin in rhesus monkeys and rats: Anatomic site and interspecies variation. Toxicol. Environ. Health 23, 207–216. Smith, J. G. (1988). Paraquat poisoning by skin absorption: A review. Hum. Toxicol. 7, 15–19. Snodgrass, H. L. (1992). Permethrin transfer from treated cloth to the skin surface: Potential for exposure in humans. J. Toxicol. Environ. Health 35, 91–105. Sreeramulu, K., Liu, R., and Sharom, F. J. (2007). Interaction of insecti­ cides with mammalian P-glycoprotein and their effect on its transport function. Biochim. Biophys. Acta 1768(7), 1750–1757. Takahashi, H., Kato, A., Yamashita, E., Naito, Y., Tsuda, S., and Shirasu, Y. (1987). Potentiations of N-methylcarbamate toxicities by organo­ phosphorous insecticides in male mice. Fundam. Appl. Toxicol. 8, 139–146. Timchalk, C., Dryzga, M. D., Langvardt, P. W., Kastl, P. E., and Osborne, D.W. (1990). Determination of the effect of tridiphane on the pharmacokinetics of [14C]-atrazine following oral administration to male Fischer 344 rats. Toxicology 61, 27–40. Timchalk, C., Nolan, R. J., Mendrala, A. L., Dittenber, D. A., Brzak, K. A., and Mattsson, J. L. (2002). A physiologically based pharmacokinetic and pharmacodynamic (PBPK/PD) model for the organophosphate insecticide chlorpyrifos in rats and humans. Toxicol. Sci. 66(1), 34–53. Turner, J. C., and Shanks, V. (1980). Absorption of some organochlo­ rine compounds by the rat small intestine – in vivo. Bull. Environ. Contam. Toxicol. 24, 652–655.

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Wester, R. C., Bucks, D. A., and Maibach, H. I. (1994). Human in vivo percutaneous absorption of pyrethrin and piperonyl butoxide. Food Chem. Toxicol. 32, 51–53. Wester, R. C., Melendres, J., Logan, F., Hui, X., Maibach, H. I., Wade, M., and Huang, K. C. (1996a). Percutaneous absorption of 2,4-dichlorophenoxyacetic acid from soil with respect to soil load and skin contact time: In vivo absorption in rhesus monkey and in vitro absorption in human skin. J. Toxicol. Environ. Health 47, 335–344. Wester, R. C., Quan, D., and Maibach, H. I. (1996b). In vitro percutane­ ous absorption of model compounds glyphosate and malathion from cotton fabric into and through human skin. Food Chem. Toxicol. 34, 731–735. Wester, R. C., Melendres, J., Sedik, L., Maibach, H. I., and Riviere, J. E. (1998). Percutaneous absorption of salicylic acid, theophylline, 2,4dimethylamine, diethyl hexyl phthalic acid and p-aminobenzoic acid in isolated perfused porcine skin flap compared to man in vivo. Toxicol. Appl. Pharmacol. 1, 159–165. Williams, P. L., Thompson, D., Qiao, G. L., Monteiro-Riviere, N. A., Baynes, R. E., and Riviere, J. E. (1996). The use of mechanistically defined chemical mixtures (MDCM) to assess component effects on the percutaneous absorption and cutaneous disposition of topically exposed chemicals. II. Development of a general dermatopharmaco­ kinetic model for use in risk assessment. Toxicol. Appl. Pharmacol. 141, 487–496. Windheuser, J. J., Haslam, J. L., Caldwell, L., and Shaffer, R. D. (1982). The use of N,N-diethyl-m-toluamide to enhance dermal and transder­ mal delivery of drugs. J. Pharm. Sci. 71, 1211–1213. Wu, H. X., Evreux-Gros, C., and Descotes, J. (1996). Diazinon toxico­ kinetics, tissue distribution and anticholinesterase activity in the rat. Biomed. Environ. Sci. 9, 359–369. Zhao, Y. H., Abraham, M. H., Le, J., Hersey, A., Luscombe, C. N., Beck, G., Sherborne, B., and Cooper, I. (2002). Rate-limited steps of human oral absorption and QSAR studies. Pharm. Res. 19(10), 1446–1457. Zhou, S. F. (2008). Structure, function and regulation of P-glycoprotein and its clinical relevance in drug disposition. Xenobiotica 38(7–8), 802–832.

Chapter 38

Metabolism of Pesticides Ernest Hodgson North Carolina State University, Raleigh, North Carolina

38.1  Introduction The word metabolism may be used to designate the sum of chemical reactions that serve to maintain life. Parts of this integrated whole are spoken of as protein metabolism, fat metabolism, nucleic acid metabolism, and the like. Such aspects, as they deal with the processing of the normal endogenous constituents of the body and the effect of pesticides on them, are dealt with in other chapters. The word metabolism may also be used to designate the effect of an organism, through its enzymes, on the chemical structure of foreign compounds now more often referred to as xenobiotics. These effects, also called biotransformation, are the subject of this chapter as they apply to pesticides. The enzymes involved in these biotransformations are frequently referred to as xenobiotic-metabolizing enzymes, XMEs. Given the enormous literature on pesticide metabolism, it is no longer possible to provide an exhaustive review of the subject. The more recent reviews and book chapters referred to throughout are recommended as sources of more detailed and recent information. It should also be noted that pesticides are not only substrates for XMEs, but may also act as inhibitors or inducers, in either case often with selectivity for specific isoforms. Inhibition and/or induction and interactions consequent to them are considered in Chapter 40.

38.2  External transformation The finding of a derivative of a compound in the tissues or excreta of an animal is not necessarily proof that the compound is the result of biotransformation in that organism. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

Compounds, especially in thin films, may undergo chemical change when exposed to light or heat. As early as 1961 Mitchell reported the effects of ultraviolet light on 141 pesticides and Matsumura (1975, 1985) summarized the effects of light and other physical factors on pesticides and their movement in the environment through 1985. The rate and extent of photochemical degradation of pesticides depends upon the chemical nature of the pesticide, the wavelength of the light, and the presence of other chemicals. The latter may act as photosensitizers, forming reactive light-energized intermediates that react with pesticides, or they may react with photoenergized pesticides. The four best known types of photochemical reactions of aromatic pesticides are ring substitution, hydrolysis, oxidation, and polymerization. Examples summarized by Matsumura (1975) include the following: substitution of a ring chlorine in 2,4-D by a hydroxyl group, hydrolysis of carbaryl, oxidation of parathion, and polymerization of pentachlorophenol. A more recent review (Stangroom et al., 2000) discusses the photochemical and thermochemical transformation in water and soil of a number of pesticides, including carbamate, organophosphorus OPs and pyrethroid insecticides and urea, chlorophenoxy, and triazine herbicides. The reactions involved are often pH-dependent and some may be catalyzed by metal and other ions. A more narrowly focused review (Pehkonen and Zhang, 2002) concerns the degradation of OPs in natural waters. The enzymes of plants and microorganisms are responsible for a wide range of biotransformations (Matsumura, 1985) and, as a result, an animal feeding on plants may ingest one or more derivatives produced in or on the plant and thus outside of the body of the animal, as well as the compound originally applied to the plant. In addition, some 893

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pesticides may be metabolized by bacteria in the intestine, still external to the cells of the animal itself. An example of the latter is reduction of parathion to aminoparathion by rumen bacteria (Ahmed et al., 1958). Although all of these possibilities should be kept in mind, they are usually of secondary importance. Moreover, many derivatives known to be formed by light, plants, or microorganisms are also formed by mammalian enzymes.

38.3  Biotransformation The most frequently studied bioconversions of pesticides are those resulting from metabolism catalyzed by the cytochrome P450 CYP-dependent monooxygenases (Ecobichon, 2001; Hodgson and Levi, 2001; Hodgson and Meyer, 1997, 2009; Hodgson et al., 1991; Kulkarni and Hodgson, 1980, 1984a,b; Kulkarni et al., 1984). The substantial number of known CYP-catalyzed reactions with pesticide substrates demonstrates the extensive knowledge base for this large group of enzymes (Hodgson and Levi, 2001). Metabolism by flavin-containing monooxygenase FMO isoforms is also frequently studied and many FMO substrates are common to both CYPs and FMOs (Hodgson and Levi, 1992; Tynes et al., 1985a,b). Metabolism by other enzymes, including phase I reactions catalyzed by prostaglandin H ­ synthetase/

cyclooxygenase COX1/2, molybdenum hydroxylases/ aldehyde AOX and xanthine oxidases, alcohol and aldehyde dehydrogenases, and esterases is often important for specific pesticides. Phase II conjugations important for pesticide metabolism include those important for the products of other xenobiotic oxidations, especially those catalyzed by the N-acetyl-, sulfo-, uridine diphosphate UDPglucuronyl-, methyl- and amino acid taurine, glycine transferases (Cerrara and Periquet, 1991). Glutathione S-transferases GSTs are important catalysts in primary reactions involving nucleophilic substitution of the many chlorinated pesticides as evidenced by the frequent detection of their mercapturates as human urinary metabolites. Depending upon the substrate, examples of both detoxication and activation can be found with any of these enzymes, although metabolic activations by CYP isoforms to form damaging electrophiles reactive with critical nucleophilic sites of proteins and DNA have been the best characterized activation reactions (Table 38.1). Human polymorphisms have been found for most, if not all, of the metabolic enzymes important for pesticide metabolism and epidemiological associations between pesticides and susceptibility to various pesticide health effects have provided information about the functional consequences of metabolism by affected enzymes. A summary of some of the methods used to demonstrate XMEs in in vitro studies is shown in Table 38.2.

Table 38.1  Metabolism of Pesticides Catalyzed by Microsomal Enzymes Reaction

Example

Reference Oxidation CYP-dependent oxidations

N-Dealkylation Nicotine→nornicotine

Papadopoulos (1964)

4-Nitrophenyl N, N-dimethyl carbamate→4-nitrophenyl N-methyl carbamate

Hodgson and Casida (1961); Strother (1972)

Dimethoate→des N-methyl derivatives

Lucier and Menzer (1970)

Dicrotophos→des N-methyl deriviatives

Tseng and Menzer (1974)

Atrazine, simazine, terbutryn→N-dealkylated derivatives

Adams et al. (1990); Rodriguez and Harkin (1995)

Diuron→desmethyldiuron

Abass et al. (2007)

Methoxychlor→mono- and dihydroxy derivatives

Bikadi and Hazai (2008); Kapoor et al. (1970); Kishimoto et al. (1995); Kurihari and Oku (1991)

Propoxur→2-hydroxyphenyl N-methyl carbamate

Oonithan and Casida (1968, 1966)

Alachlor→RCH2OH→HCHO

Jacobsen et al. (1991)

O-Dealkylation Ether cleavage

Chapter | 38  Metabolism of Pesticides

895

Table 38.1  (Continued) Reaction

Example

Reference Oxidation CYP-dependent oxidations

Ester cleavage

Epoxidation to stable epoxides

Alachlor→O-dealkylated derivatives

Hodgson et al. (1998); Coleman et al. (1999, 2000)

Chlorfenvinphos→desethyl derivatives

Donninger et al. (1967, 1972); Hutson (1981)

EPN oxon→desmethyl EPN oxon

Nomeir and Dauterman (1979)

Heptachlor→heptachlor epoxide

Khan (1969)

Aldrin→dieldrin

Kulkarni and Hodgson (1984a)

Ring hydroxylation Carbaryl→4- and 5-hydroxycarbaryl via arene oxide

Dorough and Casida (1964); Hutson (1981)

Naphthalene→naphthalene epoxide

Cho et al. (2006); Jerina et al. (1968, 1970)

Dieldrin→12-hydroxydieldrin

Hutson (1976); Baldwin et al. (1972)

Methoxychlor

Dehal and Kupfer (1994)

Terbutol

Suzuki et al. (2001)

Carbaryl→N-hydroxymethyl carbaryl

Dorough and Casida (1964); Hutson (1981); Tang et al. (2002)

Butacarb→hydroxybutyl derivatives

Douch and Smith (1971a,b)

TOCP→hydroxmethyl TOCP cyclic phosphate

Eto et al. (1962)

Pyrethrins→hydroxymethyl derivatives

Casida et al. (1975–1976); Class et al. (1991)

DEF oxidation at C adjacent to S leading to dealkylation

Hur et al. (1992)

Heterocyclic ring hydroxylation

Nicotine→hydroxynicotine

Hucker et al. (1960)

Desulfuration/ dearylation

Parathion→paraoxon

Buratti et al. (2003); Davison (1995); Foxenberg et al. (2007); Kamataki and Neal (1976); Kim et al. (2005)

Diazinon→diazoxon

Buratti et al. (2003); Poet et al. (2003); Yang et al. (1969, 1971)

Azinophos-methyl

Buratti et al. (2003)

Chlorpyrifos

Buratti et al. (2003); Poet et al. (2003); Tang et al. (2001)

Fenitrothion

Levi et al. (1988)

Other OPs→oxons

Buratti and Testai (2007); Kulkarni and Hodgson (1984a)

Dehydrogenation

- and -Chlordane→dichlorochlordene

Chadwick et al. (1975); Street and Blau (1972)

Sulfoxidation

Phorate→phorate sulfoxide→phorate sulfone

Levi and Hodgson (1988)

Vamidothion→vamidothion sulfoxide

Mehmood et al. (1996)

Thiazopyr→thiazopyr sulfoxide→thiazopyr sulfone

Feng et al. (1994)

Metam-sodium

Kim et al. (1994); Smyser and Hodgson (1985); Smyser et al. (1985, 1986)

Side chain hydroxylation

(Continued )

Hayes’ Handbook of Pesticide Toxicology

896

Table 38.1  (Continued) Reaction

Example

Reference Oxidation CYP-dependent oxidations

DEF – oxidation of S adjacent to P

Hur et al. (1992)

Diallate, Triallate and Sulfallate

Hackett et al. (1993); Mair and Casida (1991)

S-oxidation of dithianes, S-oxidation of 7-N,Ndimethylamino-1,2,3,4,5-pentathiocyclooctane

Xia et al. (1995)

Aldicarb→aldicarb sulfoxide and sulfone

Perkins et al. (1999)

Methiocarb→methiocarb sulfoxide

Buronfosse et al. (1995)

FMO-dependent oxidations N-oxidation

Sulfoxidation

Oxidative desulfuration

Nicotine→nicotine N-oxide

Tynes and Hodgson (1985a,b)

Tetram→tetram N-oxide

Hajjar and Hodgson (1980, 1982)

Phorate→phorate sulfoxide

Hajjar and Hodgson (1982b); Kim et al. (1994); Levi and Hodgson (1988)

Methiocarb→methiocarb sulfoxide

Buronfosse et al. (1995); Furnes and Schlenk (2005)

Demeton-O

Furnes and Schlenk (2005)

Ethiofencarb

Furnes and Schlenk (2005)

Fonofos→fonofos oxon

Furnes and Schlenk (2005); Hajjar and Hodgson (1980, 1982a); Smyser and Hodgson (1985); Smyser et al. (1985, 1986) Reduction

Nitro reduction

Hitchcock and Murphy (1967)

Parathion→aminoparathion Dechlorination

Esaac and Matsumura (1984)

DDT→TDE Hydrolysis DDVP→desmethyl DDVP

Hodgson and Casida (1962)

Deltamethrin→3-2,2-dibromovinyl-2,2-cyclopropane carboxylic acid and 3-phenoxybenzalde

Akhtar (1984); Anand (2006); Godin et al. (2006, 2007a,b); Ross et al. (2006)

Bioresmethrin

Ross et al. (2006)

Esfenvalerate

Godin et al. (2007)

Permethrin

Crow et al. (2007); Ghiasuddin and Soderlund (1984)

Pyrethroid-like model substrates

Huang et al. (2005)

Malathion→desethyl malathion

Buratti and Testai (2005) Conjugation

Glucuronidation

Dieldrin→dieldrin glucuronide

Baldwin et al. (1972); Hutson (1976); Matthews and Matsumura (1969)

Chapter | 38  Metabolism of Pesticides

897

Table 38.1  (Continued) Reaction

Example

Reference

Methoxychlor

Hazai et al. (2004)

Carbaryl→naphthyl glucuronide

Chin et al. (1979a,b,c)

Sulfation

Carbaryl→naphthyl sulfate

Chin et al. (1979a,b,c)

Acetylation

Fluoroacetamide→fluoroacetyl CoA

Peters (1963)

Glutathione conjugation

Methyl parathion→GSH desmethyl methyl parathion  methyl GSH

Abel et al. (2004a,b); Choi et al. (2006); Hollingworth (1969)

Epoxide hydrolase Tridiphane

Magdalou and Hammock (1987)

Table 38.2  In Vitro Tests for Microsomal Xenobiotic-Metabolizing Enzymes Reaction or enzyme

Substrate

Reference

Ester hydrolysis

Methylthiobutyrate, p-nitrophenyl acetate

Heymann and Mentlein (1981)

Benzoapyrene

Gelboin and Conney (1968); Denison et al. (1983)

Ethoxyresorufin

Aitio (1978); Burke and Meyer (1974); Pohl and Fouts (1980)

1A2rat

Acetanilide

Lewandowski et al. (1990); Mitoma and Udenfriend (1962)

1A2 human

Phenacetin

Xenotech (2008)

2A6

Coumarin

Xenotech (2008)

2B1/2 rat

Pentoxyresorufin

Lubet et al. (1985, 1990)

Benzphetamine

Werringloer (1978)

Testosterone, 16 , 16  -hydroxylation

Sonderfan et al. (1987); Wood et al. (1983)

2B6 human

Buproprion

Xenotech (2008)

2C8 human

Amodiaquine

Xenotech (2008)

2C9 human

Diclofenac

Xenotech (2008)

Cytochrome P450 1A1 rat

2C19 human

Xenotech (2008)

2D6 human

Dextromethorphan

Xenotech (2008)

2E1 rat

p-nitrophenol

Koop (1986)

2E1

Chloroxazone

Xenotech (2008)

3A1 rat

Testosterone, 6 -hydroxylation

Li et al. (1995); Sonderfan et al. (1987)

3A4/5 human

Testosterone 6, -hydroxylation

Xenotech (2008)

4A1 rat, human

Lauric acid

Kinsler et al. (1988); Xenotech (2008) (Continued )

Hayes’ Handbook of Pesticide Toxicology

898

Table 38.2  (Continued) Reaction or enzyme

Substrate

Reference

Phorate

Levi and Hodgson (1988)

N,N-Dimethylaniline

Tynes and Hodgson (1985a,b)

Methimazole

Dixit and Roche (1984)

Thiobenzamide

Cashman and Hanzlik (1981)

Flavin-containing monooxygenase

38.3.1  Biotransformation in the Liver Pesticide absorption occurs through the skin as well as the respiratory and gastrointestinal tracts, with eventual disposition to the liver from all routes of exposure, the liver being the primary site of pesticide biotransformation for the purpose of facilitating clearance through excretion of water-soluble products detoxication. However, the high level of oxidative metabolism in the liver makes this organ a possible target for more toxic metabolic products activation when detoxifying, and protective mechanisms are overwhelmed. Both acute pesticide poisonings with liver involvement and liver toxicity, including cancer, have been associated with chronic pesticide exposure. Hepatotoxicity is discussed elsewhere in this handbook.

38.3.1.1  CYP and FMO Monooxygenations As illustrated in Table 38.1, CYP carries out many different monooxygenations of pesticide substrates, such as epoxidation (e.g., aldrin), N-dealkylation (e.g., atrazine), Odealkylation (e.g., chlorfenvinphos), sulfoxidation (e.g., phorate), and oxidative desulfuration (e.g., parathion) (Ecobichon, 2001; Kulkarni and Hodgson, 1980, 1984a,b; Hodgson and Meyer, 2009). Substrates for the FMO are similarly diverse, but all are soft nucleophiles, a category that includes many organic chemicals with sulfur, nitrogen, phosphorus, or selenium heteroatoms. Although CYP isoforms appear to prefer hard nucleophiles as substrates, there is considerable overlap and most, if not all, substrates for FMO are also CYP substrates. The reverse, however, is not true, since oxidations at carbon atoms are readily catalyzed by CYP but rarely, if at all, by FMO. Moreover, even when the same substrate is oxidized by both CYP and FMO, there may be differences in the rate of oxidation, in the products, or in the stereochemistry of the same product. While isoforms of both CYP and FMO are expressed in the liver, they are broadly expressed in other organs, the proportions of different isoforms varying from organ to organ. Pesticide substrates for FMO include

organophosphates such as phorate, disulfoton, and demetonO, which yield sulfoxides; the phosphonate, fonofos, which yields fonofos oxon; carbamates such as aldicarb, methiocarb, and ethiofencarb; dithiocarbamate herbicides such as sodium metham; botanical insecticides such as nicotine; and cotton defoliants such as the trivalent organophosphorus defoliant, folex (Furnes and Schlenk, 2005; Krueger and Williams, 2005; Levi and Hodgson, 1988; Smyser and Hodgson, 1985, 1986; Smyser et al., 1985; Tynes and Hodgson, 1985a,b; Venkatesh et al., 1992a). Over 7500 animal CYP isoforms in 781 gene families have been characterized across all taxa and genomic and protein sequences are known. A system of nomenclature based upon derived amino acid sequences was proposed in 1987 and entries are continuously updated http://drnelson. utmem.edu/CytochromeP450.html. Degree of similarity in sequence classifies members to a CYP numeric gene family, then letter subfamily Nelson 2006 such that individual isoforms have unique CYP number-letter-number annotations, e.g., CYP1A1. Of the 110 animal CYP families, 18 are found in vertebrates (Nelson et al., 2008). The total number of functional CYP genes in any single mammalian species is thought to range from 60 to 200 (Gonzalez, 1990). Whereas some CYP isoforms are substrate specific, those involved in xenobiotic metabolism tend to be relatively nonspecific, although substrate preferences are usually evident. FMOs, like CYPs, are located in the endoplasmic reticulum of hepatocytes and other vertebrate cells and catalyze NADPH-dependent monooxygenation of pesticides, especially those with N, S, or P heteroatoms (Cashman and Zhang, 2006; Ziegler, 2002). Early studies on the contribution of individual CYP isoforms using partially purified CYP preparations from mouse liver showed considerable variation between fractions in oxidation of pesticide substrates, in spectral binding and in inhibition by piperonyl butoxide (Beumel et al., 1985; Levi and Hodgson, 1985). Subsequently, the use of highly purified CYPs from the livers of phenobarbital and -naphthoflavone-treated mice showed that these fractions

Chapter | 38  Metabolism of Pesticides

have much higher activity toward the organophosphorus insecticides fenitrothion, parathion, and methyl parathion than did similar fractions from the livers of untreated mice, suggesting the importance of the CYP1A and CYP2B families in these oxidations. The isoforms also produced different amounts of detoxication products as compared with the more toxic oxons with CYP2Bs forming more of the oxon (Levi et al., 1988). Similar studies showed the importance of the CYP2B family in the hepatic metabolism of phorate to phorate sulfoxide (Kinsler et al., 1988, 1990). More recently, it has become clearer that CYP2B6 is one of the most important CYP isoforms in human metabolism of pesticides (Croom et al., 2008, 2009; Hodgson and Rose, 2007; Tang et al., 2001). Studies of in vitro metabolism of pesticides were, until recently, carried out on surrogate animals. During the last decade, however, due to the availability of human liver cells, cell fractions and recombinant human XMEs, there has been an increasing number of studies of human metabolism of pesticides and, in some instances, variations due to polymorphisms have been demonstrated. A summary of pesticide substrates for human hepatic XMEs is presented in Table 38.3. It is apparent that essentially all of the human xenobioticmetabolizing CYPs, as well as some other phase I enzymes have one or more pesticide substrates. A number of studies have shown the importance of both the relative amounts of different CYP or FMO isoforms present (Buratti and Testai, 2005, 2007; Buratti et al., 2002, 2003, 2007; Cashman and Zhang, 2006; Cherrington et al., 1998b; Mutch et al., 1999, 2003; Tang et al., 2001, 2002, 2004; Usmani et al., 2002; Usmani et al., 2004a,b) as well as the effect of polymorphisms on the extent of metabolism and the distribution of metabolites (Dai et al., 2001; Tang et al., 2001); additional references can be found in Hodgson (2003).

38.3.1.2  Other Phase I Enzymes Epoxide hydrolase is another phase I enzyme known to metabolize pesticides, a well-known example being the metabolism of the herbicide, tridiphane, by the epoxide hydrolase of mouse liver (Magdalou and Hammock, 1987). The role of both carboxylesterases, CYPs, and alcohol and aldehyde dehydrogenses in the hepatic metabolism of pyrethroids has recently been studied in both rodents (Anand et al., 2006a,b; Crow et al., 2007; Godin et al., 2006, 2007; Huang et al., 2005; Price et al., 2008; Ross and Crow, 2007; Ross et al., 2006); and humans (Choi et al., 2002; Crow et al., 2007; Godin et al., 2006, 2007; Huang et al., 2005; Price et al., 2008; Ross and Crow, 2007; Ross et al., 2006); as well as the role of carboxylases in the human hepatic metabolism of malathion (Buratti and Testai, 2005); see Tables 38.1 and 38.3. The phase I metabolism of methoxychlor, because of its importance as an environmental endocrine disruptor, has received considerable attention, particularly in the laboratory

899

of the late David Kupfer (e.g., Bikadi and Hazai, 2008; Hazai et al., 2004; Hu and Kupfer, 2002a,b). Other recent studies of hepatic pesticide metabolism include those on azole fungicides (Barton et al., 2006; Mazur et al., 2007), the carbamate insecticide terbutol (Suzuki et al., 2001), and the herbicide diuron (Abass et al., 2007); see Tables 38.1 and 38.3.

38.3.1.3  Phase II Enzymes Conjugation phase II reactions of pesticides are less well known than phase I, although several types of conjugation are known to involve pesticides as substrates (Dorough, 1984; Matsumura, 1985; Mehendale and Dorough, 1972; Motoyama and Dauterman, 1980); Glucuronides are important metabolites of carbamates, including banol, carbaryl, and carbofuran (Mehendale and Dorough, 1972), as well as the endocrine disruptor, methoxychlor (Hazai, 2004) and some organophosphorus and other pesticides (Hutson, 1981). Ethereal sulfates, although not important in pesticide metabolism, may be formed from the oxidative metabolites of carbaryl and carbofuran (Dorough, 1968, 1970). Glutathione S-transferase GST is important in the metabolism of organophosphorus pesticides (Abel et al., 2004; Choi et al., 2006; Motoyama and Dauterman, 1980) and halogenated herbicides such as the chloroacetanilides and chloro-S-triazines (Abel et al., 2004; Cho and Kong, 2007). The conjugation products are typically further metabolized and, in humans, excreted as urinary mercapturic acids. Interestingly, the addition of GSH molecular weight 307 to these 200- 300molecular-weight xenobiotics creates a product that exhibits a species-dependent disposition due to differences in size thresholds for biliary transport in this range. Although not strictly speaking a detoxication reaction, hepatic aliesterase, by forming a stable phosphorylated enzyme with organophosphates oxons, may serve as an inert storage protein (Chambers et al., 1990).

38.3.1.4  Comparative Aspects Limitations of time and space militate against a comparative approach to pesticide metabolism and this chapter is devoted, with only a few exceptions, to this subject as it applies to humans and to surrogate mammals used in research and risk analysis. However, effects on fish liver are of increasing importance to environmental toxicologists and have been the subjects of many recent reports. The diversity of these reports is illustrated by the following examples: Carassius auratus and alachlor (Yi et al., 2007); Gasterosteus aculeatus and prochloraz (Sanchez et al., 2008); Ictalurus punctatus and methoxychlor (James et al., 2008); Micropterus salmoides and p,p-DDE (Barberm et al., 2007); Oncorhyncus mykiss and dieldrin (Barnhill et al., 2003); Oreochromis mossambicus and monocrotophos (Rao, 2006); Oreochromis niloticus and paraquat

Hayes’ Handbook of Pesticide Toxicology

900

Table 38.3  Human Phase I Xenobiotic-Metabolizing Enzymes Active in Pesticide Metabolisma Phase I isoform

Substrate

Reference

Alcohol dehydrogenase 

Permethrin metabolite: phenoxybenzyl alcohol

Choi et al. (2002)

Alcohol dehydrogenase -I

Permethrin metabolite: phenoxybenzyl alcohol

Choi et al. (2002)

Alcohol dehydrogenase -II

Permethrin metabolite: phenoxybenzyl alcohol

Choi et al. (2002)

Alcohol dehydrogenase 

Permethrin metabolite: phenoxybenzyl alcohol

Choi et al. (2002)

ALDH3A1

Permethrin metabolite: phenoxybenzyl aldehyde

Choi et al. (2002)

CYP1A1

Insecticides:

Aldehyde dehydrogenase

Carbaryl

Tang et al. (2002)

Carbofuran

Usmani et al. (2004b)

Sulprofos

Usmani et al. (2004b)

Herbicides: Ametryne

Lang et al. (1996, 1997)

Atrazine

Lang et al. (1996, 1997)

Terbuthylazine

Lang et al. (1996, 1997)

Terbutryne

Lang et al. (1996, 1997)

Diuron

Abass et al. (2007)

Insect repellent Deet CYP1A2

Umani et al. (2002)

Insecticides: Azinphos-methyl

Buratti et al. (2002, 2003)

Carbaryl

Tang et al. (2002)

Carbofuran

Usmani et al. (2004a)

Chlorpyrifos

Buratti et al. (2002, 2003) Foxenberg et al. (2007); Tang et al. (2001)

Diazinon

Buratti et al. (2002, 2003)

Disulfoton

Usmani et al. (2004b)

Imidacloprid

Schultz-Jander and Casida (2002)

Methiocarb

Usmani et al. (2004b)

Parathion

Buratti et al. (2003); Butler and Murray (1997); Foxenberg et al. (2007); Mutch et al. (1999, 2003); Sams et al. (2000)

Phorate

Hodgson et al. (1998)

Sulprofos

Usmani et al. (2004b)

Chapter | 38  Metabolism of Pesticides

901

Table 38.3  (Continued) Phase I isoform

Substrate

Reference

Methoxychlor

Hu and Kupfer (2002a,b)

Herbicides:

CYP2A6

Ametryne

Lang et al. (1996, 1997)

Atrazine

Lang et al. (1996, 1997)

Terbuthylazine

Lang et al. (1996, 1997)

Terbutryne

Lang et al. (1996, 1997)

Diuron

Abass et al. (2007)

Insecticides: Carbaryl

Tang et al. (2002)

Imidachloprid

Schultz-Jander and Casida (2002)

methoxychlor

Hu and Kupfer (2002)

Insect repellent: DEET CYP2B6

Usmani et al. (2002)

Insecticides: Azinophos-methyl

Buratti et al. (2002, 2003)

Carbaryl

Tang et al. (2002)

Chlorpyrifos

Buratti et al. (2002, 2003); Foxenberg et al. (2007); Tang et al. (2001)

Diazinon

Buratti et al. (2002, 2003)

Disulfoton

Usmani et al. (2004b)

Endosulfan

Casabar et al. (2006)

Imidacloprid

Schultz-Jander and Casida (2002)

Methiocarb

Usmani et al. (2004b)

Parathion

Buratti et al. (2003); Butler and Murray (1997); Foxenberg et al. (2007); Mutch et al. (1999, 2003); Sams et al. (2000)

Phorate

Usmani et al. (2004b)

Insect repellent: Deet

Usmani et al. (2002)

Herbicides: Acetachlor

Coleman et al. (2000)

Alachlor

Coleman et al. (2000)

Ametryne

Lang et al. (1996, 1997) (Continued )

Hayes’ Handbook of Pesticide Toxicology

902

Table 38.3  (Continued) Phase I isoform

CYP2C8

Substrate

Reference

Atrazine

Lang et al. (1996, 1997)

Butachlor

Lang et al. (1996, 1997)

Metolachlor

Lang et al. (1996, 1997)

Terbutryne

Lang et al. (1996, 1997)

Insecticides: Carbaryl

Tang et al. (2002)

Carbofuran

Usmani et al. (2004a)

Parathion

Mutch et al. (2003)

Phorate

Hodgson et al. (1998)

Deltamethrin, esfenvalerate

Godin et al. (2007a,b)

Methoxychlor

Hu and Kupfer (2002)

Herbicide: Ametryne CYP2C9

Lang et al. (1997)

Insecticides: Chlorpyrifos

Foxenberg et al. (2007); Tang et al. (2001)

Imidacloprid

Schultz-Jander and Casida (2002)

Methoxychlor

Bikadi and Hazai (2008)

Parathion

Foxenberg et al. (2007)

Phorate

Hodgson et al. (1998)

Herbicide: Ametryne CYP2C9*1

CYP2C9*2

Lang et al. (1997)

Insecticides: Carbaryl

Tang et al. (2002)

Disulfoton

Usmani et al. (2004b)

Methiocarb

Usmani et al. (2004b)

Phorate

Usmani et al. (2004b)

Sulprofos

Usmani et al. (2004b)

Methoxychlor

Hu and Kupfer (2002)

Insecticides: Carbaryl

Tang et al. (2002)

Chapter | 38  Metabolism of Pesticides

903

Table 38.3  (Continued) Phase I isoform

CYP2C9*3

CYP2C18

CYP2C19

Substrate

Reference

Disulfoton

Usmani et al. (2004b)

Sulprofos

Usmani et al. (2004b)

Insecticides: Carbaryl

Tang et al. (2002)

Sulprofos

Usmani et al. (2004b)

Insecticides: Carbaryl

Tang et al. (2002)

Disulfoton

Usmani et al. (2004b)

Phorate

Usmani et al. (2004b)

Sulprofos

Usmani et al. (2004b)

Insecticides: Azinphos-methyl

Buratti et al. (2002)

Carbaryl

Tang et al. (2002)

Carbofuran

Usmani et al. (2004a)

Chlorpyrifos

Buratti et al. (2002); Foxenberg et al. (2007); Tang et al. (2001)

Deltamethrin

Godin et al. (2007)

Esfenvalerate

Godin et al. (2007)

Diazinon

Buratti et al. (2002); Kappers et al. (2001)

Disulfoton

Usmani et al. (2004b)

Fipronil

Tang et al. (2004)

Imidacloprid

Schultz-Jander and Casida (2002)

Methiocarb

Usmani et al. (2004b)

Parathion

Buratti et al. (2002); Foxenberg et al. (2007); Mutch et al. (2003)

Phorate

Hodgson et al. (1998)

Sulprofos

Usmani et al. (2004b)

Deltamethrin

Godin et al. (2007)

Esfenvalerate

Godin et al. (2007)

Methoxychlor

Bikadi and Hazai (2008); Hu and Kupfer (2002)

Insect repellent: (Continued  )

Hayes’ Handbook of Pesticide Toxicology

904

Table 38.3  (Continued) Phase I isoform

Substrate

Reference

Deet

Usmani et al. (2002)

Herbicides:

CYP2C19*1B

Ametryne

Lang et al. (1996, 1997)

Atrazine

Lang et al. (1996, 1997)

Terbuthylazine

Lang et al. (1996, 1997)

Diuron

Abass et al. (2007)

Insecticide: Chlorpyrifos

CYP2C19*8

Insecticide: Chlorpyrifos

CYP2C19*6

Tang et al. (2001)

Insecticide: Chlorpyrifos

CYP2D*1

Tang et al. (2001)

Insecticide: Chlorpyrifos

CYP2C19*5

Tang et al. (2001)

Tang et al. (2001)

Insecticides: Carbaryl

Tang et al. (2002)

Carbofuran

Usmani et al. (2004b)

Disulfoton

Usmani et al. (2004b)

Sulprofos

Usmani et al. (2004b)

Methoxychlor

Hu and Kupfer (2002)

Insect repellent: Deet CYP2D6

Usmani et al. (2002)

Insecticides: Diazinon

Sams et al. (2000)

Imidacloprid

Schultz-Jander and Casida (2002)

Methiocarb

Usmani et al. (2004b)

Parathion

Mutch et al. (2003)

Herbicide: Atrazine CYP2E1

Insecticides:

Lang et al. (1997)

Chapter | 38  Metabolism of Pesticides

905

Table 38.3  (Continued) Phase I isoform

Substrate

Reference

Carbaryl

Tang et al. (2002)

Imidacloprid

Schultz-Jander and Casida (2002)

Parathion

Mutch et al. (2003)

Phorate

Hodgson et al. (1998); Usmani et al. (2004b)

Insect repellent: Deet

Usmani et al. (2002)

Herbicide: Atrazine CYP3A4

Lang et al. (1997)

Insecticides: Azinphos-methyl

Buratti et al. (2002, 2003)

Carbaryl

Tang et al. (2002)

Carbofuran

Usmani et al. (2004a)

Chlorpyrifos

Buratti et al. (2002, 2003); Dai et al. (2001); Foxenberg et al. (2007); Tang et al. (2001)

Diazinon

Buratti et al. (2003); Kappers et al. (2001); Sams et al. (2000)

Dimethoate

Buratti and Testai (2007)

Disulfoton

Usmani et al. (2004b)

Endosulfan

Casabar et al. (2006)

Fipronil

Tang et al. (2004)

Imidacloprid

Schultz-Jander and Casida (2002)

Methiocarb

Usmani et al. (2004b)

Parathion

Buratti et al. (2003); Butler and Murray (1997); Foxenberg et al. (2007); Kappers et al. (2001); Mutch et al. (1999, 2003); Sams et al. (2000)

Phorate

Hodgson et al. (1998); Usmani et al. (2004b)

Sulprofos

Usmani et al. (2004b)

Vamidothion

Mehmood et al. (1996)

Methoxychlor

Hu and Kupfer (2002)

Insect repellent: Deet

Usmani et al. (2002)

Herbicides: Acetachlor

Coleman et al. (1999, 2000); Hodgson et al. (1998) (Continued  )

Hayes’ Handbook of Pesticide Toxicology

906

Table 38.3  (Continued) Phase I isoform

CYP3A4-F189S

Substrate

Reference

Alachlor

Coleman et al. (1999, 2000); Hodgson et al. (1998)

Ametryne

Lang et al. (1996, 1997)

Atrazine

Lang et al. (1996, 1997)

Butachlor

Lang et al. (1996, 1997)

Terbuthylazine

Lang et al. (1996, 1997)

Terbutryne

Lang et al. (1996, 1997)

Diuron

Abass et al. (2007)

Insecticide: Chlorpyrifos

CYP3A4-L293P

Insecticide: Chlorpyrifos

CYP3A4-M445T

Dai et al. (2001)

Insecticide: Chlorpyrifos

CYP3A5

Dai et al. (2001)

Insecticide: Chlorpyrifos

CYP3A4-P467S

Dai et al. (2001)

Dai et al. (2001)

Insecticide: Carbaryl

Tang et al. (2002)

Carbofuran

Usmani et al. (2004a)

Chlorpyrifos

Foxenberg et al. (2007)

Parathion

Mutch et al. (2003); Foxenberg et al. (2007)

Phorate

Mutch et al. (2003); Usmani et al. (2004b)

Sulprofos

Usmani et al. (2004b)

Deltamethrin

Godin et al. (2007)

Esfenvalerate

Godin et al. (2007)

Methoxychlor

Hu and Kupfer (2002)

Insect repellent: Deet CYP3A7

Usmani et al. (2002)

Insecticides: Carbofuran

Usmani et al. (2004a)

Chlorpyrifos

Foxenberg et al. (2007)

Chapter | 38  Metabolism of Pesticides

907

Table 38.3  (Continued) Phase I isoform

FMO1

FMO3

Carboxylase

Substrate

Reference

Parathion

Foxenberg et al. (2007)

Insecticides: Aldicarb

Schlenk et al. (2002)

Demeton-O

Furnes and Schlenk (2005)

Disulfoton

Usmani et al. (2004b)

Ethiofencarb

Furnes and Schlenk (2005)

Fonofos

Furnes and Schlenk (2005)

Methiocarb

Furnes and Schlenk (2005); Usmani et al. (2004b)

Sulprofos

Usmani et al. (2004b)

Phorate

Hodgson et al. (1998); Usmani et al. (2004b)

Insecticide: Aldicarb

Schlenk et al. (2002)

Demeton-O

Furnes and Schlenk (2005)

Ethiofencarb

Furnes and Schlenk (2005)

Fonofos

Furnes and Schlenk (2005)

Insecticides: Bioresmethrin

Ross et al. (2006)

Permethrin

Crow et al. (2007); Ross et al. (2006)

Pyrethroid-like model substrates

Huang et al. (2005)

Malathion

Buratti and Testai (2005)

a

Adapted from Hodgson 2003.

(Figueiredo-Fernandes et al., 2006); and Rhamdia quelen and glyphosate (Glusczak et al., 2007); Salmo salar and p,p-DDE (Mortensen and Arukwe, 2006). There is a much smaller body of literature on birds (e.g., Cortright and Craigmill, 2006), reptiles (e.g., Gunderson et al., 2006), and various food and feral mammals (e.g., Dupuy et al., 2001).

38.3.2  Biotransformation in Extrahepatic Tissues The liver is generally more important than other organs in the biotransformation of xenobiotics, including pesticides. However, other organs and tissues may be active to some degree. For example, it was shown early that DDT

is degraded by rat diaphragm, kidney, and brain in vitro (Judah, 1949). Later study showed that these changes proceeded at a very slow rate in vivo. However, as shown in the following sections, not all extrahepatic metabolism is inefficient. Some enzymes outside the liver may be induced, but the matter has received little attention. Wattenberg (1971) demonstrated that the small intestines of rats fed a balanced purified diet or starved for 1 day possess virtually no benzo[a]pyrene hydroxylase activity, whereas the intestines of rats fed the same diet plus turnip greens, broccoli, cabbage, or brussels sprouts have marked activity of this enzyme. The same activity in human skin is induced by polycyclic hydrocarbons (Alvares et al., 1973). Neal (1972) showed that monooxygenases of the lung active in the

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metabolism of parathion can be induced by phenobarbital. Induction of enzymes metabolizing pesticides is considered in detail in Chapter 40.

38.3.2.1  Lung The lung is a primary site of exposure to airborne as well as blood-borne environmental pollutants, such as pesticides, and for this reason is a target organ for many chemically induced toxicities (Bond, 1983, 1993; Dahl and Lewis, 1993; Ding and Kamienski, 2003). Because the lung has a full complement of metabolic enzymes, it has the capacity to activate and deactivate pesticides and other xenobiotics. Early studies showed that parathion is metabolized to paraoxon and diethylphosphorothioic acid by rabbit lung at about 20% of the rate in liver (Neal, 1972). Several studies have demonstrated the importance of pulmonary CYP and FMO enzymes in pesticide oxidation (Feng et al., 1990; Li et al., 1992). In the lung, FMO appears to play a more important role than CYP in the oxidation of certain pesticides and xenobiotics (Kinsler et al., 1988; Tynes and Hodgson, 1983, 1985a,b). Other studies have shown the existence of an FMO form now known as FMO2 in the lung not present in the liver (Lawton et al., 1990; Tynes and Hodgson, 1983; Tynes et al., 1985; Venkatesh et al., 1992b; Williams et al., 1984, 1985). Boland et al. (2004) demonstrated the metabolism of naphthalene, primarily to the dihydrodiol, in respiratory tissues of Rhesus monkeys and the metabolism and toxic effects of naphthalene in respiratory tissues continues to be of interest (Bogen et al., 2008).

38.3.2.2  Nasal Tissues The nasal mucosa is the first tissue of contact for inhaled xenobiotics and compounds have been identified that cause nasal lesions or tumors in experimental animals. The drug-metabolizing activity of nasal tissues has been reviewed by Reed (1993) and, more recently, by Ding and Kamienski (2003). Enzymes known to be present include a variety of CYPs CYP1A1, 2B1, 2E1, 3A1, 4A1, 2G1, FMOs, carboxylesterases, epoxide hydrolases, glutathione s-transferases, and uridine 5-diphosphate UDP glucuronyl transferases. It is of some interest that, despite the low concentrations of nasal CYP enzymes, these have been demonstrated to have greater specific activity toward several substrates than liver CYPs; perhaps as a result of higher ratios of NADPH cytochrome P450 reductase to CYP in the nasal tissues. Nasal CYPs appear to be less inducible than liver isoforms, although they appear to be sensitive to a number of CYP inhibitors. Few pesticides are known to give rise to toxic endpoints in the nasal tissues. However, alachlor, a restricteduse chloroacetamide herbicide that at one time was widely used in agriculture, was demonstrated to cause rare nasal carcinomas in rats. The putative metabolic product thought

to be responsible for its carcinogenicity was identified as diethylbenzoquinoneimine DEBQI, which is produced only after extensive metabolism of alachlor, involving CYPs as well as an aryl amidase. Human CYP isoforms 2B6 and 3A4 are among those that have been identified as being important in the production of metabolite precursors to DEBQI (Coleman et al., 1999, 2000). Genter and co-workers (Deamer et al., 1994; Genter et al., 1995, 1998) have demonstrated the role of ­microsomal epoxide hydrolase and CYP2E1 in the nasal toxicity of dichlobenil in the mouse. It was subsequently shown that CYP2A10 and 2A11, isoforms that comprise some 25% of the total olfactory cytochrome CYP content, also play an important role in the nasal toxicity of dichlobenil (Ding et al., 1994, 1996).

38.3.2.3  Skin Because the skin is the largest organ in the human body, is continuous over the surface area of the body and is in direct contact with the environment, it is often the portal of entry for pesticides and other xenobiotics. The skin is known to contain many of the xenobiotic metabolizing enzymes found in the liver, and some of these have been shown to be inducible, primarily by polycyclic hydrocarbons (Goerz et al., 1994; Jugert et al., 1994; Baron et al., 2008). The metabolic capacity of skin for pesticides was shown early when slices of rabbit skin were shown to hy­­drolyze paraoxon at a concentration of 7.7  103 M to the extent of 20% in 1 h/g of tissue. Because absorption of paraoxon and related compounds is slow, this metabolism may be an important defense mechanism (Fredriksson et al., 1961). By use of in vitro methods, such as the isolated perfused porcine skin flap (Carver et al., 1990) and mouse skin microsomes (Venkatesh et al., 1992a), the skin has been shown to have the capacity to metabolize a variety of pesticides. For example, Chang et al. (1994), using the isolated perfused porcine skin flap, showed that both carbaryl and parathion were metabolized during uptake by the skin.

38.3.2.4  Kidney Because of the kidney’s high blood flow, its ability to concentrate chemicals, and the presence of renal xenobiotic metabolizing enzymes, the kidney may also be a site of toxicity from xenobiotics. Many of these toxic effects can be directly attributable to the presence and localization of specific forms of enzymes responsible for activation (Hu et al., 1993; Speerschneider and Dekant, 1995). Several studies have highlighted the importance of renal oxidative enzymes, particularly FMO, in the metabolism of pesticides and other xenobiotics (Kinsler et al., 1988; Tynes and Hodgson, 1983). As was the case with the lung, the renal FMO enzymes played a greater role in microsomal systems in the oxidation of several pesticides than renal CYP, suggesting an important

Chapter | 38  Metabolism of Pesticides

role for FMO in the extrahepatic metabolism of toxicants. Studies of kidney FMO have provided evidence for several isoforms in the kidney, including the forms found in liver and lung (Atta-Asafo-Adjei et al., 1993; Burnett et al., 1994; Ripp et al., 1999; Venkatesh et al., 1991) and Furnes and Schlenk (2005) have demonstrated the sulfoxidation of fenthion and methiocarb by kidney FMO.

38.3.2.5  Central Nervous System Very little is known about XMEs in the central nervous system CNS. Several studies have demonstrated CYP activity and constitutive expression of various CYP isozymes (Britto and Wedlund, 1992; Ghersi-Egea et al., 1993; Hansson et al., 1992; Hodgson et al., 1993; Miksys and Tyndale, 2009). The activation or detoxication of pesticides by the CNS is of particular interest in the case of pesticides that both exhibit their action and are metabolized in the brain. Studies by Chambers and Chambers (1989) demonstrated that the neurotoxicity of a series of organophosphorus compounds correlated better with activation in the brain than with activation in the liver. Several studies have reported activity of known FMO substrates by brain microsomes (Bhamre et al., 1993; Duffel and Gillespie, 1984; Kawaji et al., 1994), and one form of FMO has been demonstrated using polymerase chain reaction PCR amplification (Blake et al., 1996).

38.3.2.6  Gastrointestinal Tract Xenobiotic metabolism in the gastrointestinal tract has been reviewed by Ding and Kaminski (2003). Some carbaryl is hydrolyzed and the resulting naphthol is conjugated with glucuronic acid by the intestine (Pekas and Paulson, 1970) and recently Furnes and Schlenk (2005) have demonstrated sulfoxidation of fenthion by FMO in the intestine.

38.4  Toxicity of metabolites In general, metabolites are less toxic than their parent compounds, if for no other reason than that they are usually more water-soluble and, therefore, more rapidly excreted. There are notable exceptions for which biotransformation results in an inherently more toxic product. Such reactions are generally referred to as activation reactions. These reactive metabolites may combine covalently with cellular constituents such as DNA, RNA, or protein and carcinogenesis, mutagenesis, and cellular necrosis are often attributable to such reactive metabolites (Anders et al., 1992; Guengerich, 1992, 1993; Levi and Hodgson, 2001; Parke, 1987). Hollingworth et al. (1995) reviewed reactive metabolites with particular reference to agrochemicals, The metabolic production of a more toxic compound is sometimes called lethal synthesis to emphasize that biotransformation in this instance is the source of danger.

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The term lethal synthesis was introduced in a lecture given on June 7, 1951, by Peters (1952) in connection with fluoroacetic acid. This compound is not itself an enzyme inhibitor, but is converted by enzymes into a highly toxic material. Peters (1963) later reviewed and extended the concept of lethal synthesis, although the term itself is now seldom used. The effects of metabolism may be complex, as illustrated by studies of bromobenzene. It has been known for some time that the liver necrosis associated with this compound is caused by one or more toxic metabolites. Stimulation of its biotransformation by phenobarbital potentiates the injury of toxic doses to the liver, and inhibition of its metabolism by SKF 525-A prevents this injury. However, although 3-methylcholanthrene causes a slight in vitro stimulation of the metabolism of bromobenzene and does not alter the overall rate in vivo, it does protect against the hepatotoxicity. Rats dosed with ­ bromobenzene after induction with 3-methylcholanthrene excrete more bromophenyldihydrodiol, bromocatechol, and 2-­bromophenol than do uninduced rats. Increase in the first two compounds suggests an increased capacity to detoxify the highly reactive epoxide. Increase in 2-bromophenol suggests that induction by 3-methylcholanthrene diverts the metabolism of bromobenzene to a comparatively nontoxic pathway (Zampaglione et al., 1973). Of particular concern has been the role of metabolic activation in the carcinogenic process, particularly in the formation of DNA adducts by reactive metabolites. For example, monooxygenase enzymes have been postulated to play a role in the metabolic activation of alachlor and metolachlor (Brown et al., 1988; Feng and Wratten, 1989; Feng et al., 1990; Jacobsen et al., 1991; Li et al., 1992). Although studies suggest that alachlor has a greater carcinogenic potential than metolachlor, the carcinogenic response to these compounds are species- and tissue-specific, alachlor being a nasal-specific carcinogen in rats but not in mice. Metolachlor, on the other hand, is carcinogenic to the liver but not to nasal tissue U.S. Environmental Protection Agency, 1986, 1987. Available evidence suggests that the species- and tissue-specific responses observed, particularly for alachlor, result from specific metabolic enzymes, including monooxygenases and arylamidases, and the generation of the putative carcinogenic metabolite, diethylbenzoquinone imine (see Coleman et al., 1999, 2000 for appropriate references). The most studied generation of reactive metabolites from pesticides is the generation of oxons from organophosphorus compounds containing the P  S moiety, by oxidative desulfuration. Not only does this reaction produce the oxons, cholinesterase inhibitors responsible for the neurotoxicity of these compounds, but it also releases reactive sulfur, a potent CYP inhibitor (Neal, 1980; Neal and Halpert, 1982; Neal et al., 1983); see Chapter 40 for further information on inhibition. The hepatic metabolism

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of organophosphorus insecticides continues to be investigated in both humans and rodents. In general, through a common intermediate containing a phosphiithrane ring, they are either activated to their oxons, potent inhibitors of acetylcholinesterase and other esterases, or detoxified. This intermediate is generated by CYP isoforms, particularly CYP1A2; CYP2B6, and CYP 3A4 (see Tables 38.1 and 38.3). The mode of action of the insecticide synergist, piperonyl butoxide, and other methylenedioxyphenyl compounds is also due to a reactive metabolite, believed to be a carbene derivative, that combines with the heme iron of CYP (Dahl and Hodgson, 1979; see Chapter 40).

38.5  Physiological factors affecting biotransformation Species, strains, and individuals may all vary in their susceptibility to toxicants, including pesticides. In some cases it has been possible to explain these differences by one of several causes, including differences in metabolism. In this section, some examples are presented in which it is known that the activity of microsomal enzymes is influenced by age, gender, and species.

38.5.1  Developmental Effects Microsomal XME activity is low in the fetus and the newborn, but increases rapidly during the early days or weeks of life (Croom et al., 2009; Fouts and Adamson, 1959; Ronis and Cunny, 1994, 2000). For this reason fetuses and newborns are often more susceptible to certain drugs and xenobiotics than adults. With aging, there is generally a decrease in enzymatic activity, although increases in some activities have been observed (Kitahara et al., 1982; Van Bezooijen, 1984; Van Bezooijen et al., 1986). In the mouse, FMO1 and FMO5 are expressed as early as gestation days 15 and 17 and equally between genders until puberty. FMO3 is not expressed until 2 weeks postpartum and is found equally in male and female until 6 weeks postpartum, when it becomes undetectable in the male. This developmental pattern, as seen in the female mouse, is similar to that seen in humans of either gender (Cherrington et al., 1998a).

38.5.2  Species Differences Many species differences in the metabolism of xenobiotics can be explained in terms of differences in activity of liver microsomal enzymes (Brodie and Maickel, 1962; Quinn et al., 1958; Walker, 1994), although broad similarities often exist across large systematic groups. For example, Barron et al. (1993) observed that the detoxication of chlorpyrifos in

Hayes’ Handbook of Pesticide Toxicology

channel catfish was similar to that in other vertebrates in both phase I and phase II metabolism. An example of a comparative study of a specific pesticide is that of Chin et al. (1979b) on carbaryl. The gender-dependent expression of FMO isoforms also varies between species, as outlined below.

38.5.3  Individual and Strain Differences Strain and individual differences often are discussed conveniently in terms of tolerance and resistance, both implying reduced susceptibility to a toxicant. The word tolerance is used when the observed decrease in susceptibility occurs in an individual organism as a result of its own previous or continuing exposure to the particular toxicant or to some other conditioning stimulus while resistance refers to a change in a population brought about by genetic selection. Not only may differences in basic levels of enzyme activity be detected in different species, but this has also been seen in different strains of mice (Jay, 1955), rats (Quinn et al., 1958), rabbits (Cram et al., 1965), and birds (Ronis and Walker, 1989; Walker, 1983). Comparative aspects of xenobiotic metabolism, particularly as they relate to CYP, have been reviewed by Hodgson (1979) and Kulkarni et al. (1975).

38.5.4  Gender Differences Metabolism of xenobiotics may vary with the gender of the organism and in some cases differences in overall toxicity between males and females of various species are known (Bonate, 1991). In the absence of induction, microsomal enzyme activity is often higher in the adult male rat than in females or immature males. However, the stimulatory effect of xenobiotics on microsomal enzymes is usually greater in females and immature males than in the adult male rat (Conney and Burns, 1962). Gender differences become apparent at puberty and are usually maintained throughout adult life. The differences in microsomal monooxy­genase activity between males and females have been shown to be under the control of sex hormones, at least in some species. Sexually dimorphic CYPs appear to arise, in the rat, by programming, or imprinting, that occurs in neonatal development. This imprinting is brought about by a surge of testosterone that occurs in the male neonate and appears to imprint the developing hypothalamus so that in later development growth hormone is secreted in a gender-specific manner. This pattern of growth hormone production pulsatile and the higher level of circulating testosterone in the male maintain the expression of male-specific CYP isoforms such as 2C11. On the other hand, a more continuous pattern of growth hormone secretion and the lack of circulating testosterone appear to be responsible for female-specific CYPs such as 2C12 (Gonzalez, 1989; Hosteter et al., 1987; Kobliakov et al., 1991; Schenkman et al., 1989).

Chapter | 38  Metabolism of Pesticides

Gender-specific expression is seen also with the FMO enzymes. It has been known for some time that hepatic FMO activity is higher in female mice than males and that the lower levels in males result from testosterone repression (Dannan et al., 1986; Duffel et al., 1981; Falls et al., 1997; Lemoine et al., 1991; Wirth and Thorgeirsson, 1978). In addition, hormonal changes during pregnancy have been reported to increase FMO levels (Williams et al., 1985). With regard to pesticide oxidation, gender differences have also been observed, with higher activity in female mouse liver than in male (Kinsler et al., 1988). Recent studies have identified FMO isozymes involved in these gender differences and some of the hormonal factors involved in regulation. In several strains of mouse liver, FMO1 expression was found to be two to three times higher in female mice compared to males, and FMO3, expressed in females at levels comparable to FMO1, was not detectable in male liver (Falls et al., 1995, 1997; Cherrington et al., 1998a). In rat liver, however, FMO1 is higher in the male, whereas FMO3 is gender-independent in both rat and human. FMO5 is gender-independent for mouse, rat, and human (Cherrington et al., 1998a).

38.5.5  Genetic Factors The existence of discontinuous or biphasic variation is a strong indication of the possibility of genetic involvement. Some examples are given here in which such variation in enzyme activity within populations has been proved to be genetic in origin. The fungicide ziram caused hemolytic anemia with Heinz body formation in a man later shown to be deficient in erythrocyte G-6-PD. Ziram also caused one of the typical in vitro reactions formation of Heinz bodies in the blood of another person known to be deficient in this enzyme (Pinkhas et al., 1963). Many of the phenotypic variations in drug responses observed in human populations have been shown to result from polymorphisms in the expression of the xenobiotic metabolizing enzymes. (For reviews see Smith et al., 1994a,b; Kalow, 1991; Coutts and Urichuk, 1999; and Wormhaudt et al., 1999.) Many human CYP isoforms have been shown to be polymorphic (Daly et al., 1998; Goldstein and De Morais, 1994; Smith et al., 1994a,b). Pesticides metabolized by these polymorphic CYP isoforms such as chlorpyrifos by CYP2B6 and CYP3A4 could have higher or lower risk factors in some proportion of the exposed population dependent upon the distribution of the relevant polymorphisms (Dai et al., 2001; Hodgson and Rose, 2007; Tang et al., 2001).

38.6  Tolerance and resistance The terms resistance and tolerance refer to a relative lack of susceptibility of a population of organisms to the effects

911

of a toxicant. If the genetic trait preexists in a population so that it is obvious when the population is first exposed to the toxicant, this should be regarded as tolerance rather than resistance, the latter term being reserved for those cases in which the trait is brought to an observable level only through selection brought about by exposure to the toxicant. It is clear that many instances of resistance are based on differences in toxicant metabolism that distinguish the resistant from the nonresistant population.

38.6.1  Tolerance Tolerance to a compound is often the result of an organism’s increased ability to metabolize the chemical subsequent to an initial exposure. This is true, for example, in connection with the pesticides nicotine (Werle and Uschold, 1948) and dieldrin (Wright et al., 1972). In a few instances, it has been shown clearly that the increased metabolism responsible for tolerance was mediated by higher activity of the microsomal enzymes of the liver. It seems likely that the same explanation will hold in connection with some other instances of tolerance. As recorded in Chapter 40 pesticides frequently act as in-ducers of microsomal enzymes. Because activity of these enzymes usually leads to detoxication, it seems likely that many of the compounds listed as inducers are capable of producing tolerance under suitable conditions. Tolerance also may exist in situations in which it has been impossible to demonstrate any increased ability to metabolize the toxicant; finally, there are instances of tolerance for which the mechanism is not only unknown but unexplored. For example, rodents may develop true tolerance as distinguished from bait shyness to a number of rodenticides including arsenic oxide, zinc phosphide, strychnine, sodium fluoroacetate, ANTU, and norbormide (Lund, 1964, 1967). Certain populations of pine mice subjected to control with endrin lost susceptibility to the compound, but sublethal exposure conferred a degree of tolerance regardless of the past history of the population (Webb and Horsfall, 1967).

38.6.2  Resistance Resistance in the toxicological sense is better known in insects and a variety of other pest species than in vertebrates. Many species with public health importance have long been known to be resistant to one or more pesticides (Brown and Pal, 1971; Georghiou and Saito, 1983) and a much larger number of agricultural pests are also resistant. The numbers in both groups continue to grow. Resistance to a particular compound does not involve an entire species but only the toxicant-stressed population; nevertheless, resistance constitutes a serious public health and economic problem.

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Resistant strains are first recognized only after the parent population has been selected by killing off many of its susceptible members. The rate at which selection progresses depends not only on the intensity of the selection pressure, but also on the duration of each generation. Therefore, it is not surprising that resistance has most often been observed among organisms such as bacteria or houseflies, characterized by large numbers of individuals and a rapid rate of reproduction. However, resistance has also been observed in species with relatively small populations and relatively slow multiplication. Resistance of a vertebrate species to a pesticide apparently was first recognized in the 1960s and 1970s in connection with Norway rats exposed to warfarin. This phenomenon was first reported from Scotland (Boyle, 1960) and subsequently from Denmark (Lund, 1964, 1967), England and Wales (Bentley, 1969; Drummond, 1966), the Netherlands (Ophof and Langeveld, 1969), Germany (Telle, 1971), and the United States, specifically North Carolina and Idaho (Brothers, 1972; Jackson and Kaukeinen, 1972). Resistance in rats continues to be reported and investigated on a regular basis, for example in France (Lasseur et al., 2005), in Germany (Pelz, 2007), and in China (Wang et al., 2008). The early literature on the resistance of mammals to warfarin was reviewed by Lund (1967). Only a few points need to be recorded here. So far, resistance is known to occur in four species, the Norway and roof rats, the house mouse, and humans. In addition, in their original studies of coumarin compounds, Link and his students reported marked variation in susceptibility in rabbits as a Mendelian characteristic (Campbell et al., 1941). The exact mechanism of inheritance of resistance to warfarin is not clear. In humans, the facts are consistent with transmission by a single autosomal dominant gene (O’Reilly et al., 1963), but this was based on only one kindred. In rats and mice, it seems that more than one gene is involved. The physiological basis of the resistance in vertebrates also is not clear and may be different in different instances. It is generally held that it involves the vitamin K cycle (Cain et al., 1998) and polymorphisms in the VKOR vitamin K epoxide reductase gene (Lasseur et al., 2005; Wang et al., 2008). The role of CYP isoforms in anticoagulant resistance is not clear, although it has been noted in bromadiolone resistance in the rat, several CYP isoforms are overexpressed (Markussen et al., 2008a,b). In every instance studied, including humans, resistance extended to other coumarin anticoagulants and those based on indanedione, while susceptibility to heparin is normal. Another early report of resistance among vertebrates involved mosquito fish collected from insecticide-contaminated waters near cotton fields (Vinson et al., 1963). Further study revealed two- to 500-fold levels of resistance to a variety of pesticides in mosquito fish and five other species of fish (Boyd and Ferguson, 1964a,b; Ferguson and Bingham, 1966a,b; Ferguson and Boyd, 1964; Ferguson

Hayes’ Handbook of Pesticide Toxicology

et al., 1964, 1965). Resistance to chlorinated hydrocarbon pesticides was found in three species of frog (Boyd et al., 1963). The degree of resistance may be so great in some instances that resistant species can withstand enough poison to kill their predators. Ozburn and Morrison (1962) were the first to produce resistance to a pesticide in a mammal by selection under laboratory conditions. In mice selected by a single intraperitoneal dose of DDT administered at 4 weeks of age, resistance in the ninth generation had increased by a factor of 1.7 as measured by the LD50. Although the factor of 1.7 is small, about half of the susceptible mice withstood a dose that was uniformly fatal to control mice. Further study (Ozburn and Morrison, 1965) revealed that the selected and control colonies differed in their rates of oxygen consumption. The resistant mice were fatter than the susceptible ones and considerable evidence indicated that resistance depended on preferential deposit of DDT in the fat and consequently the avoidance of peak levels in sensitive tissues. The resistance was not specific for DDT but extended to lindane and dieldrin (Barker and Morrison, 1966). Success in the development of resistance in mammals in the laboratory has not been uniform; apparently some strains are not sufficiently heterozygous to respond to selection (Guthrie et al., 1971). Thus many instances are known in which species or strains differ in their susceptibility to pesticides, the resistance arising through selection in the field. In other instances, it has been possible to produce resistance in the laboratory through selection. In many instances it has been possible to define the genetic mechanisms responsible for observed differences in the metabolism of the pesticides in question. Except in the case of insects, a genetic mechanism has been defined only rarely in connection with metabolism of pesticides. The following references are suggested on this subject: Dauterman (1994), Evered and Collins (1984), Georghiou and Saito (1983), and Hayes et al. (1990).

Conclusions Knowledge of the metabolism of pesticides is essential for several reasons, including the development of more selective insecticides, and provides, in part, the fundamental basis for science-based risk assessments for human and environmental health (Buratti et al., 2007; Hodgson and Rose, 2005, 2007). Until recently, and as a matter of necessity, this research was carried out almost exclusively on experimental animals and the results, particularly in the case of human health risk assessments, extrapolated to humans. Although much essential background will continue to be obtained from experimental animals, due to the ready availability of human hepatocytes, human cell lines, human cell fractions, and recombinant human enzymes,

Chapter | 38  Metabolism of Pesticides

essential information will, in the future, be derived to a greater extent directly from human studies. At the same time, studies utilizing surrogate animals will also be revolutionized by such new techniques of molecular biology as the use of knockout and transgenic including “humanized” mice (e.g., Gonzalez, 2003), and the knowledge of the genomes of many species. These same techniques through the study of genetic polymorphisms will enable us to identify human populations at increased risk and enable comparative studies to be carried out at the level of specific isoforms of the XMEs involved. Thus the study of pesticide metabolism continues to evolve into a new, more molecular, era that will be fascinating as well as useful.

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Chapter | 38  Metabolism of Pesticides

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Chapter | 38  Metabolism of Pesticides

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Chapter 39

Distribution and Pharmacokinetics* Kelly J. Dix Research Triangle Institute, Research Triangle Park, North Carolina

Pesticides and other chemicals interact with biological systems, including people, in many ways. Pesticides are inherently toxic. To understand how nontarget species’ (e.g., humans’) exposure to a pesticide may result in tox­ icity, one must first understand how the pesticide enters, moves about in, and is eliminated from the body. More specifically, one must characterize the processes of absorp­ tion of the chemical after various routes of exposure (e.g., dermal, inhalation, oral), distribution of the chemical to different organs and tissues in the body, metabol­ism of the chemical, and elimination of the chemical and/or its metabolites from the body. Pharmacokinetics, the study of the absorption, distribution, metabolism, and elimi­ nation of chemicals, utilizes mathematical models to describe the time course of the chemical in the body. This chapter includes a general discussion of distribution and pharmacokinetics.

39.1  Introduction For a pesticide to elicit toxicity, it must be transferred from the external site of exposure to the target site (e.g., organ, nucleic acid, receptor) and achieve a sufficiently high con­ centration in the target organ (Figure 39.1). Absorption is the translocation of the pesticide from an external source of exposure to the bloodstream. Once in the blood, the chemical is distributed through the body and delivered to tissues, where it may leave the blood and enter the cells of the tissue or remain in the blood and simply pass through the tissue. In certain tissues, such as the liver, the chemical may be effectively removed from the body by metabolism. Other tissues, such as kidney and lung, serve to eliminate xenobiotics from the body by excretion. Absorption, distri­ bution, metabolism, and excretion, which are collectively

*Editorially abbreviated and reformatted from Dix, Kelly J. “Absorption, Distribution and Pharmacokinetics,” Chapter 24 in the Handbook of Pesticide Toxicology (Krieger, R., Editor), 2nd edition, 2001. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

Oral

Dermal

Gastrointestinal tract

Skin

iv, ip, sc, im

Inhalation

Lung

Excretion

Excretion Liver

Blood

Metabolites Kidney

Target organ(s)

Excretion

Figure 39.1  Representation of the absorption, distribution, metabol­ ism, and excretion of toxicants.

termed disposition, are all factors that affect the concen­ tration of a chemical in target tissues. Pharmacokinetics refers to the mathematical description of the time course of chemical disposition in the body. Metabolism and excretion are discussed in detail in other chapters of this work. This chapter focuses on absorption, distribution, and pharmacokinetics.

39.2  Distribution Once in the bloodstream, the chemical is available for distribution throughout and elimination from the body. Metabolism and excretion, which are components of elimi­ nation, are discussed in other chapters. This section will focus on distribution, the reversible translocation of chemi­ cals from one location to another in the body. Distribution of a toxicant to and accumulation in the target organ may result in toxicity. Accumulation at nontarget sites, on the other hand, results in storage of the pesticide away from the site of action and ultimately protection from toxicity. The physiology of the organism and the physicochemical characteristics of the pesticide are important factors in the distribution of absorbed pesticides. 923

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39.2.1  Total Body Water Chemicals in the body move throughout the water com­ partments of the body. As already discussed, the ability of chemicals to move between the various water compartments is limited by the physicochemical properties of the chemi­ cal. Total body water consists of plasma water, interstitial water, and intracellular water. In humans, approximately 60% of body weight is water, with plasma, intracellular, and interstitial water accounting for 5, 15, and 40% of body weight, respectively. Plasma water, which represents approximately 53–58% of blood volume in humans, plays an essential role in the distribution of absorbed chemicals. For a chemical to move from blood (plasma water) into tissues, it must cross the endothelial cell layer lining the capillaries (i.e., capillary wall) to enter the interstitial water, then cross the plasma membrane to enter the intra­ cellular water. Chemicals exist in blood as free circulating chemicals or are noncovalently bound to plasma proteins. The rates of association and dissociation with plasma pro­ teins are very rapid (on the order of milliseconds), and it is assumed that the bound and free forms of the chemical are in equilibrium. The capillary wall is permeable to small molecules, but not readily permeable to high molecular weight molecules such as plasma proteins. Only free chem­ icals that are small enough to pass through the capillaries, then, are available to move from plasma water to intersti­ tial water. The processes for crossing plasma membranes described in Chapter 37 govern passage from the intersti­ tial water to intracellular water.

Initial distribution is influenced primarily by blood flow to tissues, whereas final distribution is influenced primar­ ily by the relative affinity of the chemical for various tissues relative to blood (i.e., the tissue partition coefficient). In the early phase of distribution, tissues that receive a high blood flow (e.g., liver, kidney, and brain) may achieve high con­ centrations of the chemical even though the tissue partition coefficient for that chemical is low. Likewise, tissues that are slowly perfused (e.g., adipose) may achieve a low concentra­ tion of the chemical in the early phase of distribution even though the tissue partition coefficient for that chemical is very high. Later in the distribution phase, however, the chem­ ical redistributes to tissues based on tissue partition coeffi­ cients, and the chemical is more concentrated in tissues with relatively high partition coefficients. Pesticides and other xenobiotics do not have the same tissue partition coefficient for all tissues. For example, dimethoate has a relatively high affinity for liver, muscle, and brain (Garcia-Repetto et al., 1995), whereas the chlorinated insecticides DDT, aldrin, and dieldrin are lipophilic and have high affinities for adipose tis­ sue (Lehman, 1956; Robinson et al., 1969).

39.2.3  Volume of Distribution When a chemical is absorbed and distribution is complete, its concentration in blood depends on the amount absorbed and the extent of tissue distribution. The apparent volume of distribution (Vd) is a proportionality constant that relates the amount of chemical in the body to its concentration in plasma,

39.2.2  Rate and Extent of Distribution Factors that influence the rate and extent of distribution of a chemical to a particular tissue include blood flow to the tissue (rate of delivery), the mass of the tissue, the ability of the chemical to cross membranes, and the affinity of the chemical for the tissue relative to blood. The rate of dis­ tribution of a chemical from blood to tissues can be per­ fusion- or diffusion-rate-limited. For lipophilic chemicals that rapidly cross membranes, the rate of delivery to tissues is limited by blood flow (perfusion-rate-limited). For polar and ionized chemicals that do not readily cross the plasma membrane, the rate of delivery to tissues is limited by dif­ fusion (diffusion-rate-limited). Plasma protein binding increases the rate of distribution to tissues for toxicants that are not diffusion-rate-limited. The free toxicant may read­ ily cross the capillary wall, effectively decreasing its free concentration in blood. Bound toxicant then dissociates from plasma proteins to maintain the equilibrium between the bound and free forms, yet the new free molecules rap­ idly leave the blood, which further increases dissociation of bound toxicant, and so on. In contrast, distribution of more polar compounds that are diffusion-rate-limited is dependent on the extent of protein binding.



Vd 

amount in body concentration in plasma



where Vd is the theoretical volume of fluid the chemical would occupy to achieve the observed concentration in plasma and does not necessarily correspond to the volume of a particular body fluid compartment. For example, a chemical that is sequestered in a particular tissue will have a low concentration in plasma and a corresponding high volume of distribution, which may in fact be greater than the total body water.

39.2.4  Blood–Brain Barrier The blood–brain barrier, which protects the cen­ tral nervous system, is not an absolute barrier. 2,4Dichlorophenoxyacetic acid (2,4-D), for example, has been measured in the rabbit brain after an intravenous (IV) dose (Kim et al., 1996). The tight junctions of the capillary endothelial cells and the surrounding glial cell processes are the main structural features that contribute to the low permeability of the blood–brain barrier. Chemicals that cir­ culate in blood must pass through the capillary endothelial

Chapter | 39  Distribution and Pharmacokinetics

cell membrane and the glial cell membrane to access the interstitial fluid of the brain. The low protein content of the brain’s interstitial fluid limits lipophilic chemicals that are highly bound to plasma proteins. Chemical access to the brain, then, is limited to those that are free (unbound), lipo­ philic, nonionized, and transported by specialized carrier systems, whereas ionized and highly plasma protein bound chemicals are excluded by the blood–brain barrier. Another barrier to brain access is the presence of an adenosine 5triphosphate (ATP)-dependent multidrug-resistance (MDR) protein, which transports intracellular chemicals back into the extracellular space. Since the blood–brain barrier is not fully developed at birth, the risk of toxicity from exposure to some chemicals is higher for newborns and young chil­ dren than for adults.

39.2.5  Placental Transfer Functions of the placenta include delivery of nutrients to the fetus, removal of fetal waste, and maternal/fetal blood gas exchange, which suggests that many chemicals move freely across the placental membrane. In the framework of distribution, the placenta is not a barrier to protect the fetus from exposure to toxicants. Rather, the placenta is a typical plasma membrane barrier that is permeable to lipophilic and nonionized molecules that readily cross plasma membranes, thereby exposing the fetus to toxicants. 1,1-Dichloro-2,2-bis(p-chlorophenyl)ethylene (DDE) and 2,4-D have been demonstrated to cross the placenta in rats, rabbits, and bats (Kim et al., 1996; Sandberg et al., 1996; Thies and McBee, 1994), and prenatal human exposure to 2,4-D has been associated with mental retardation in off­ spring (Casey and Collie, 1984).

39.2.6  Storage and Redistribution Chemicals may accumulate in body compartments due to protein binding, active transport processes, or high solu­ bility in (i.e., affinity for) a particular tissue. These sites of accumulation can be considered storage depots. Whereas a chemical in any tissue compartment is in equilibrium with its free concentration in blood, storage is dynamic. Removal of free chemical from the body by metabolism or excretion shifts the equilibrium such that stored chemical is released.

39.2.6.1  Plasma Proteins Plasma protein binding plays a very important role in chemical-induced toxicity. Displacement of one chemical from plasma proteins by another chemical can have severe consequences. If the bound chemical is very toxic, its dis­ placement results in a higher free concentration in plasma, which results in greater availability for distribution to its site of toxic action.

925

39.2.6.2  Fat Adipose is a storage depot for a number of highly lipophilic chemicals, including pesticides. Storage in adipose tissue may be considered a protective mechanism in that the pes­ ticide is stored in a nontarget tissue, thereby lowering its concentration at the site of toxic action. For example, the chlorinated insecticides DDT (Dale et al., 1962; Hayes et al., 1958), chlordane (Ambrose et al., 1953), hexachloro­ benzene isomers (Davidow and Frawley, 1951), lindane (Ludwig et al., 1964), aldrin, and dieldrin (Robinson et al., 1969) are lipophilic and accumulate in fat. Upon dieting and starvation, fat is mobilized and the stored chemical is released, which results in a sudden increase in the blood concentration of the pesticide and availability of the pes­ ticide for redistribution. As was described for plasma pro­ teins, chemicals stored in adipose tissue may be displaced by other chemicals. Street (1964) demonstrated that DDT displaces dieldrin from its storage sites in rat adipose, yet methoxychlor does not affect dieldrin storage. Toxicity may be observed if the released chemical is redistributed to the target organ.

39.2.6.3  Other Tissues and Tissue Components Sequestration in tissues (e.g., kidney and liver) may be due to interaction of chemicals with tissue macromolecules such as proteins and nucleic acids, which influence the affinity of a tissue for a given chemical (tissue partition coefficient). Bone tissue, for example, is a potential stor­ age depot for heavy metals.

39.2.7  Storage with Repeated Exposure Body burden is the term for the concentration (or amount) of chemical in the body at any given time, and the biologi­ cal half-life of a chemical is the time required to reduce the concentration of the chemical in the body by onehalf, in the absence of further intake. Many pesticides are water soluble and easily excreted or readily metabolized to more water-soluble compounds that are easily excreted. Lipophilic pesticides such as the organochlorines, however, are stored in fat and are not easily removed from the body, and most people around the world carry a low body burden of organochlorine pesticides (Burgaz et al., 1994; Durham, 1969; Zatz, 1972). Repeated exposure to a chemical may result in cumulative storage and an increased body burden of the chemical. If the interval between exposures is long relative to the biological half-life of the chemical, all or most of the chemical will be removed from the body prior to subsequent exposures, and it is unlikely that the chemi­ cal will accumulate in the body. If the interval between exposures is short relative to the biological half-life, on the other hand, there will be a residual body burden from the first exposure when the second exposure occurs, and so on, such that the chemical accumulates in the body.

Hayes’ Handbook of Pesticide Toxicology

926

Cumulative storage of a chemical upon repeated expo­ sure continues until a steady state of storage is reached. Factors that influence storage include exposure level (dos­ age), time interval between exposures, duration of repeated exposures, interaction with other chemicals, age, sex, spe­ cies, disease status, and nutritional status. See below for a mathematical discussion of storage.

39.3  Pharmacokinetics

39.3.1  Noncompartmental Models Noncompartmental models use statistical moment theory for analysis of plasma concentration–time data. Area under the plasma concentration vs. time curve (AUC) is a mea­ sure of the total systemic exposure to the chemical. AUC is the integral of the rate of change of concentration in plasma as a function of time: AUC 



Pharmacokinetics is the modeling and mathematical description of the time course of chemical disposition (absorption, distribution, metabolism, and excretion). Although urine and exhaled breath may be obtained from humans, blood is the only tissue that can be readily and repeatedly sampled in humans. Pharmacokinetic models typically describe the change in blood (or plasma) con­ centration of the chemical with time. There are two basic approaches to characterizing the pharmacokinetics of a chemical in the body: compartmental and noncompartmen­ tal. Compartmental pharmacokinetic models represent the body as discrete compartments with mathematical descrip­ tions of the movement of chemical between compartments, including the processes of absorption and elimination. Compartmental models may be subdivided into classical and physiologically based models. In contrast to compart­ mental models, the noncompartmental approach assumes no compartmentalization of the body and applies the trapezoidal rule for calculating the area under the plasma concentration–time curve to characterize a chemical’s pharmacokinetics.

∞

∫0

C dt

(1)



The first moment of the plasma concentration vs. time curve is the plasma concentration multiplied by time vs. time curve, and the area under the first moment curve (AUMC) is AUMC 



∞

∫0

tC dt



In pharmacokinetic studies, plasma samples are not collected through infinite time, but rather the collection period ends at some time T. AUC and AUMC may be approximated from zero time to T using the trapezoidal rule (Figure 39.2) and then can be extrapolated from time T to  as ∞

∫T

∞

∫T



CT 

(3)

TCT C  T2  

(4)

C dt 

tC dt 

Approximated AUC = Area1 + Area2 + Area3 + Area4 + (CT/β) C1

Concentration

Area1 = (C1+C0)(t1–t0)/2 Area2 = (C2+C1)(t2–t1)/2

C2

Area3 = (C3+C2)(t3–t2)/2 Area4 = (C4+C3)(t4–t3)/2

C3

Extrapolated Area =CT/β

C4 = CT C0

t0

t1

t2

(2)

t4 = T

t3 Time

Figure 39.2  Determination of area under the plasma concentration vs. time curve using the trapezoidal rule (refer to Section 39.3.1).

Chapter | 39  Distribution and Pharmacokinetics

927

In Eqs. (3) and (4), CT is the observed concentration at the last time point T, and  is the slope of the terminal elimination phase of the log plasma concentration vs. time curve. The mean residence time (MRT), which represents the time that is required for 63.2% of the chemical to be eliminated, is



MRT 

AUMC AUC

(5)

Bioavailability (F), which is the fraction of chemi­ cal that is absorbed after extravascular administration, may also be calculated using noncompartmental models. Because the process of absorption is bypassed with intra­ venous administration, bioavailability after intravenous administration is assumed to be unity. Bioavailability is determined for extravascular administration (e.g., oral, der­ mal) with reference to an intravenous dose as F

Div AUCex Dex AUCiv



(6)

D Cl  iv AUC

(7)

Cl can be calculated after an extravascular dose only if the bioavailability is 100% (i.e., F  1). The apparent volume of distribution at steady state (Vss can be calculated after a single IV dose) is



Vss  Cl  MRTiv

(8)

The first-order elimination rate constant (ke) and elimi­ nation half-life (t1/2) can also be calculated after a single

1 ke

MRTiv 

(9)



t1/ 2  0.693MRTiv

(10)

For chemicals that cannot be described by a one-com­ partment model,





where Div and Dex are the intravenous and extravascu­ lar doses, and AUCiv and AUCex are the areas under the plasma concentration vs. time curve after intravenous and extravascular doses. When the intravenous and extravascu­ lar doses are the same, F is simply the proportion of AUC after extravascular and intravenous doses. Bioavailability is often referred to as a percentage. For example, the bio­ availability of orally administered permethrin is 0.61 or 61% (Anadon et al., 1991). A number of factors, including route of administration and species, may affect bioavail­ ability. For example, the bioavailability of paraquat was 45, 12, or 3.8% after an intratracheal, oral, or dermal dose, respectively (Chui et al., 1988). Species-dependent bio­ availability has been shown for orally administered meto­ sulam, which was only 20% bioavailable in mice and dogs, but greater than 70% in rats (Timchalk et al., 1996). Plasma clearance (Cl), a measure of the inherent abil­ ity to remove a chemical from the body, is the volume of plasma that is cleared of the chemical per unit time. Cl after an intravenous dose is calculated as



IV dose for chemicals that appear to be characterized by a one-compartment model (see Section 39.3.3) from the relationships

1 ke′

MRTiv 

ke′ 

Cl Vss

(11)



(12)



and the effective elimination half-life is the product of 0.693 and MRTiv. Repeated exposure to a chemical at constant time inter­ vals may lead to accumulation of the chemical in the body until a steady state is achieved. During any exposure inter­ val  at steady state, the rate of chemical entry into the body is equal to the rate of its elimination (i.e., amount absorbed equals amount eliminated). Wagner (1967) pro­ posed the concept of a concentration index (RC), which provides information with regard to the increased accumu­ lation with multiple exposures. RC is defined as the ratio of the average concentration of a chemical in blood during an exposure interval of length  at steady state ( C∞ ) and the average concentration in blood during the same time inter­ val after a single exposure ( C ),



RC 

C∞

(13)

C

where C∞ and C are defined as





C∞ 

1

 ∫t

t2

C∞ dt

1

C

1  C dt  ∫0



(14)

(15)

In Eq. (14), C is the concentration of the chemical in blood or plasma at time t after dosing during steady state, and   t2  t1. When C and C are measured, C∞ and C can be calculated from the respective concentration vs. time curves using the trapezoidal rule.

Hayes’ Handbook of Pesticide Toxicology

928

39.3.2  Overview of Classical Compartmental Models Classical compartmental models typically divide the body into one or more compartments that have no physiologi­ cal or anatomical reality (Figure 39.3). It is assumed that the rate of transfer between compartments, as well as the rate of elimination from the compartments, are linear or first-order processes. Each model has an associated series of mathematical equations that describe the absorption, elimination, and transfer of chemicals between compart­ ments. These equations are dependent only on the model structure and are independent of the chemical under study. Classical compartmental models can provide important parameters that describe chemical disposition, including the volume of distribution, absorption and elimination rate constants, elimination half-life, and plasma clearance. In the following discussion, elimination is assumed to occur only from compartment 1, which is referred to as the cen­ tral compartment.

39.3.3  One-Compartment Model A one-compartment model (Figure 39.4), which repre­ sents the body as a single homogeneous compartment, ade­ quately describes the pharmacokinetics of chemicals that

rapidly equilibrate between blood and tissues. Therefore, it is reasonable to assume that the concentration of the chem­ ical in blood (or plasma) is proportional to its concentra­ tion at the site of toxicity.

39.3.3.1  Intravenous Bolus Dose In the simplest one-compartment model, the chemical is introduced directly into the single compartment, and elimi­ nation occurs by a first-order process (Figure 39.4). The single compartment has a volume Vd, which in this case is the apparent volume of distribution. A typical plasma concentration–time curve for a one-compartment system is shown in Figure 39.4. This system is mathematically described by a first-order equation in which the rate of removal of the chemical (mass per time) is proportional to the body load of the chemical (mass), dA  ke A dt



where A is the amount of chemical in the body (units of mass) and ke is the first-order elimination rate constant (units of reciprocal time), which represents the fractional elimina­ tion of chemical per unit time. A solution to Eq. (16) is At  A0 exp(ke t )



Compartment 1

Compartment 1

Compartment 2

Compartment 3

(16)

Compartment 2

Compartment 1

Compartment 1

Figure 39.3  Schematic representations of one-, two-, and three-compartment models.

Compartment 2

(17)

Chapter | 39  Distribution and Pharmacokinetics

929

iv Dose, A0

Central compartment volume = Vd

100

8

Concentration

Concentration

10

ke

6 4 2

10

y-intercept = C0 = A0/Vd slope = -ke/2.303 C(t)

1

1/2 C(t)

0.1

t1/2

0.01

0

Time

Time

Figure 39.4  Representation of a one-compartment model with IV administration and first-order elimination, including a typical plasma concentra­ tion vs. time profile (linear and logarithmic scales). The volume of distribution, elimination rate constant, and elimination half-life are estimated by graphical methods.

where At is the amount of chemical in the body at time t, and A0 is the amount of chemical in the body at time zero. More frequently, concentration rather than the amount of chemical is measured in plasma, and Eq. (17) is rewritten as

Ct  C0 exp(ke t )

(18)

where Ct and C0 are the concentrations (units of mass/vol­ ume) of the chemical in plasma at time t and time zero, respectively. Taking the logarithm of both sides of Eq. (18) yields



log Ct  log C0 

ke t 2.303





(19)

(20)

The apparent volume of distribution is the volume into which the initial dose (A0) would have to be dissolved to achieve the initial concentration of the chemical in plasma, C0.

(21)

Equation (21) can be solved for t1/2, t1/ 2 

0.693 ke

(22)



which may also be estimated by inspection of the graph of log Ct vs. t (Figure 39.4). Plasma clearance (Cl) is a measure of the inherent abil­ ity to remove a chemical from the body. Cl represents the volume of plasma that is cleared of the chemical per unit time and is the ratio of the rate of elimination (mass/time) and concentration (mass/volume): Cl 



C0  C0 exp (ke t1/ 2 ) 2



The graph of log Ct vs. t has a y intercept of C0 and a slope of ke/2.303; hence ke can be determined from the slope of the log Ct vs. t graph (Figure 39.4). The appar­ ent volume of distribution, Vd, can be determined from the known amount of chemical introduced into the body by intravenous injection at time zero and the intercept of the log Ct vs. t graph as A Vd  0 C0

The elimination half-life (t1/2) of a chemical is the time required for the amount or concentration of chemical in plasma to decrease by one-half in the absence of additional exposure. Therefore, Ct is equal to one-half of C0 after one half-life has passed since the dose was administered, and

k A k CV dA/dt  e t  e t d  keVd Ct Ct Ct



(23)

Integration of Eq. (23) yields



Cl 

dose AUC

Equation (23) can also be rearranged to solve for

(24)

Hayes’ Handbook of Pesticide Toxicology

930

ke 



Cl Vd

(25)

Substitution of Eq. (24) into Eq. (25) and rearrange­ ment leads to the equation for Vd: Vd 



dose AUC  ke





Ct 



dA  ka Aa  ke A dt

(29)

dA  ka Aa dt

(28)

Extravascular dose

ka

ka FD exp(ke t )  exp(ka t ) Vd k a  ke

(30)

A typical plasma concentration–time curve for a com­ pound that is absorbed by a first-order process rapidly equilibrates between blood and tissues, and is eliminated by a first-order process as shown in Figure 39.5. After oral administration to rats, the plasma concentration vs. time profiles for triclopyr (Timchalk et al., 1996), diazinon (Wu et al., 1996), and paraquat (Chui et al., 1988) are all described by the model in Figure 39.5. Some time after administration, absorption is essentially complete and Eq. (30) is reduced to

(27)

In Eq. (27), Aa is the mass of chemical at the site of absorption and A is the mass of chemical in the body. As was noted in Section 39.3.1, extravascular exposure to chemicals is different from intravenous exposure in that it cannot be assumed that 100% of the dose is absorbed. Some fraction F of the dose (D) is absorbed, or only the product FD is bioavailable. The rate of removal of the chemical from the site of absorption is

Ct 



ka FD exp(ke t ) Vd ka  ke

(31)

Taking the logarithm of both sides of Eq. (31) yields log Ct  log



ka FD kt  e Vd (ka  ke ) 2.303

(32)

The postabsorption phase of the graph of log Ct vs. t has a slope of ke/2.303, and as in the case of a onecompartment model with an intravenous dose, ke can be determined from the terminal slope of the log Ct vs. t graph (Figure 39.6). The absorption rate constant may be

Central compartment, volume = Vd

10

ke

100

8

Concentration

Concentration

exp(ke t )  exp(ka t ) k a  ke

which can be rewritten in terms of concentration to yield

Humans are not typically exposed to pesticides by the intravenous route, but by extravascular routes (oral, der­ mal, inhalation), and the pesticide must be absorbed to enter the blood. Absorption is assumed to occur by a firstorder process with an absorption rate constant ka as shown in Figure 39.5. For extravascular exposure, then, the rate of removal of the chemical from the body is the net difference in the rates of introduction (by absorption) and elimination (by metabolism and excretion):



At  ka FD

(26)

39.3.3.2  Extravascular Dose



Solving for A as a function of time in the preceding equations yields

6 4 2 0

10 1 0.1 0.00

Time

Time

Figure 39.5  Representation of a one-compartment model with first-order absorption (i.e., extravascular administration) and first-order elimination, including a typical plasma concentration vs. time profile (linear and logarithmic scales).

Chapter | 39  Distribution and Pharmacokinetics

931

obtained from the y intercept of the plasma concentration vs. time graph, where the intercept is kaFD/(Vd (ka  ke)), or by the method of residuals as shown in the example in Figure 39.6 and Table 39.1. Integration of Eq. (32) from zero time to infinity yields AUC 

1 ka FD 1     Vd (ka  ke )  ke ka 



(33)

1967) derived the following equation for the average plasma concentration during any interval  at steady state:



C AUC 



FD Vd ke

(34)



Cl and Vd are derived from Eqs. (25) and (26), where dose is adjusted for bioavailability:



Cl 

FD AUC

FD Vd  AUC  ke

(37)





1  exp(ke  ) 1  exp(ka  )  ka FD      Vd  (ka  ke )  ke ka

(38)

Substituting Eqs. (37) and (38) into Eq. (13) and rear­ ranging yields the concentration index RC 

(35)

1    1   ka exp(k  )  ke exp(k  ) e a    k  k  k a  ke  a e

(39)

(36)



For chemicals with ka  ke, which is the case for many of the organochlorine pesticides, for instance (Zatz, 1972),

39.3.3.3  Storage with Repeated Extravascular Exposure As discussed in Section 39.3.1, repeated administration of a given dose (D) of a chemical at fixed time intervals () eventually leads to a steady state (equilibrium) of stor­ age (Figure 39.7). The following discussion continues to assume that first-order processes govern absorption and elimination. Accumulation of a chemical in the body is described by the concentration index defined in Eq. (13). Wagner and colleagues (Wagner et al., 1965; Wagner, 10

Concentration (mg/l)

FD FD  Cl Vd ke 

The average plasma concentration during the first dose interval (from time zero to ) is

which reduces to



C∞ 

C

FD (1  exp(ke  )) Vd ke 



(40)

and RC 



1 1  exp(ke  )



(41)

The accumulation ratio (RA) for multiple exposures at fixed time intervals was defined by Wagner (1967) as the

Dose

ka

Volume = Vd

ke

y-intercept = kaFDose/Vd(ka-ke) 1 slope = -ke/2.303

0.1

slope = -ka/2.303

0.01 0

6

12 Time (h)

18

Figure 39.6  Estimation of volume of distribution and absorption and elimination rate constants for the one-compartment model in Figure 39.5 by graphical methods (i.e., curve stripping). Data are shown in Table 39.1 (refer to Section 39.3.3.2).

Hayes’ Handbook of Pesticide Toxicology

932

Substitution of Eq. (22) into Eq. (42) yields Table 39.1  Data Used for the Method of Residuals Example Shown in Figure 39.6 (One-Compartment Model with IV Administration and First-Order Elimination) Time (h)

Plasma

Extrapolated concentration (mg/l) Plasma

Residual

1.44 FDt1/ 2 

(43)

1.44 FDt1/ 2 1.44t1/ 2   FD

(44)

A∞ 

Hence,



RA 

0.25

0.218

1.083

0.865

0.5

0.382

1.057

0.675

1

0.597

1.005

0.408

39.3.4  Multicompartment Models

2

0.759

0.910

0.151

4

0.724

8

0.499

12

0.334

16

0.224

20

0.150

24

0.101

Many chemicals do not rapidly equilibrate between blood and tissues, and their plasma concentration–time profiles do not conform to the one-compartment model already described. Instead, elimination from plasma is multi­ phasic, the simplest case being biphasic elimination. The early phase is referred to as the distribution phase, and the later phase is the postdistribution or elimination phase. The plasma concentration of the chemical declines more rapidly during the distribution phase compared to the elimination phase. Two schematics of two-compartment models are shown in Figure 39.3. The central compartment includes blood and tissues in which the chemical rapidly equilibrates (e.g., tissues that receive a high blood flow) and is considered a homogeneous compartment. This is analogous to the single compartment of a one-compartment model. The peripheral compartment consists of tissues for which equilibrium is not instantaneous. In classical com­ partmental models, the chemical moves between the cen­ tral and peripheral compartments with associated transfer rate constants, but elimination is assumed to occur only from the central compartment with an associated elimina­ tion rate constant (Figure 39.8). Absorption, distribution, and elimination are assumed to be first-order processes.

Concentration

95% of steady state concentration

τ

t1/2

τ Time to 95% steady state

39.3.4.1  Intravenous Bolus Dose

Time Figure 39.7  Approach to steady state plasma concentration with repeated administration at constant dose intervals (). See Section 39.3.3.3 for calculation of pharmacokinetic parameters.

ratio of the average mass of chemical in the body during any exposure interval at equilibrium and the average mass of chemical absorbed after a single exposure. The aver­ age mass of chemical absorbed after a single exposure is simply FD, the percent of dose absorbed. Use of the relationship between concentration, mass, and volume of distribution shown in Eq. (20) allows Eq. (37) to be rear­ ranged to solve for average mass of chemical during an exposure interval at steady state:



A∞ 

FD ke 



A schematic of a two-compartment model with first-order elimination is shown in Figure 39.8. The diazinon plasma concentration vs. time after an IV dose is represented by this two-compartment model (Wu et al., 1996). The con­ centration of chemical in plasma after an intravenous bolus dose as a function of time can be expressed as the sum of two monoexponential terms,

Ct  Aet  Be t

(45)

At some time after dosing, the distribution phase is complete and the only process that contributes to removal of the chemical from plasma is elimination. During this time, Eq. (45) reduces to

(42)

Ct  Be t

(46)

Chapter | 39  Distribution and Pharmacokinetics

933

iv Dose

Central compartment

k12

Peripheral compartment

k21

k10 10

8

Concentration

Concentration

10

6 4 2 0

1

0.1

Time

Time

Figure 39.8  Representation of a two-compartment model with IV administration and first-order elimination, including a typical plasma concentra­ tion vs. time profile (linear and logarithmic scales).

where B is the intercept and  is the slope of the terminal phase of the log Ct vs. t curve (Figure 39.9). The rate con­ stant  is analogous to ke in the one-compartmental model described in Section 39.3.3. The elimination half-life is estimated from  according to the equation



0.693 

Concentration

t1/ 2 

10

(47)

The method of residuals is used to estimate A and  (Figure 39.9 and Table 39.2). The initial concentration, C0, is determined by substituting t  0 into Eq. (45):

C0  A  B

(48)

C(t) =Ae -αt + Be-βt y-intercept = A slope = -α/2.303

1

y-intercept = B slope = -β/2.303

0.1

0.01 0

6

12 Time

(49)



The rate constants  and  are composites of k12, k21, and k10 with the relationships





24

Figure 39.9  Estimation of volume of A, B, , and  for the twocompartment model in Figure 39.10 by graphical methods (i.e., curve stripping). Data are shown in Table 39.2 (refer to Section 39.3.4.1).

Similar to the one-compartment model, the volume of the central compartment is dose Vc  AB

18

k10 

 k21



k12      k21  k10

(53)

(54)



    k12  k21  k10

(50)

39.3.4.2  Extravascular Dose



a  k10 k21

(51)

A schematic of a two-compartment model with first-order absorption and elimination is shown in Figure 39.10. Absorption and elimination are assumed to occur via the central compartment only. The plasma concentration as a function of time can be expressed as

The rate constants k12, k21, and k10 are determined from the relationships below (see Gibaldi and Perrier, 1982, for derivations).



k21 

A  B AB

(52)



Ct  A exp(t )  B exp( t )  C exp(k01t ) (55)

It is often difficult to distinguish the absorption phase from the distribution phase of the log Ct vs. t curve

Hayes’ Handbook of Pesticide Toxicology

934

because k01 is similar in magnitude to . Estimation of rate constants after an extravascular dose often requires data after an intravenous dose to distinguish between k01 and  (Gibaldi and Perrier, 1982).

are connected by blood flow, and chemicals may enter the body by any route. The model in Figure 39.11 incorporates exposure by the oral, dermal, and inhalation routes and elimination by urinary excretion, exhalation, and metabo­ lism. PBPK models use mathematical descriptions of chemical disposition that are based on the physiological, physicochemical, and biochemical determinants of dispo­ sition, which include biochemical reaction rates and tissue partition coefficients for the chemical and the physiology (e.g., organ volumes, blood flows, respiration rates) of the animal. PBPK models do not assume that all processes governing disposition are linear, and saturable metabolism (Michaelis–Menten kinetics), for example, is easily incor­ porated into PBPK models. PBPK models exist for pesti­ cides from a variety of chemical classes (Table 39.3). Like classical compartmental models, the compart­ ments in Figure 39.11 represent organs or tissue groups in which a chemical is uniformly distributed, and arrows represent the pathways that govern chemical disposition. Each compartment has an associated volume. Absorption is represented by arrows to the portals of entry for various routes of exposure, blood flow is represented by arrows that interconnect the compartments of the model, and metabolism and excretion are represented by arrows from the compartments in which these processes occur. PBPK models require three types of parameters as inputs: physio­ logical, physicochemical, and biochemical. Physiological parameters (e.g., pulmonary ventilation rate, cardiac out­ put, blood flow to tissues, and tissue volumes) for humans and several laboratory animal species are available in the literature (Arms and Travis, 1988). Physiological parame­ ters, which are not dependent on the chemical under study, are assumed to be constant for a given species. However, if

Table 39.2  Data Used for the Method of Residuals Example Shown in Figure 39.9 (Two-Compartment Model with IV Administration and First-Order Elimination) Time (h)

Plasma

Extrapolated concentration (mg/l) Plasma

Residual

1

6.534

0.488

6.046

2

4.318

0.468

3.850

4

1.988

0.431

1.558

8

0.614

0.364

0.250

12

0.345

0.308

0.037

16

0.264

20

0.220

24

0.187

39.3.5  Physiologically Based Models Unlike classical compartmental models, physiologically based pharmacokinetic (PBPK) models represent physiologi­ cal and anatomical reality (Figure 39.11). The compartments

Extravascular dose

k01

k12

Central compartment

k21

Peripheral compartment

k10

10

8

Concentration

Concentration

10

6 4 2

1

0

0

Time

Time

Figure 39.10  Representation of a two-compartment model with first-order absorption (i.e., extravascular administration) and first-order elimination, including a typical plasma concentration vs. time profile (linear and logarithmic scales).

Chapter | 39  Distribution and Pharmacokinetics

935

Inhaled dose

Exhaled Alveolar space Lung blood Kidney

Excreted

Poorly perfused tissues Fat Rapidly perfused tissues

Dermal dose

Skin Liver

Oral dose

Metabolized Figure 39.11  Schematic representation of a physiologically based pharmacokinetic model. This model contains descriptions of exposure by the inhalation, oral, and dermal routes, and elimination by exhalation, excretion, and metabolism (refer to Section 39.3.5).

Table 39.3  Some Existing PBPK Models for Pesticides from Various Chemical Classes Pesticide

References

Dieldrin

Leung and Paustenbach (1988)

Kepone (chlordecone)

el-Masri et al. (1995, 1996); Yang et al. (1995a,b)

Lindane (hexachlorocyclohexane)

DeJongh and Blaauboer (1997)

Hexachlorobenzene

Freeman et al. (1989); Roth et al. (1993)

Diisopropylfluorophosphate

Gearhart et al. (1994)

Dichlorobenzene

Hissink et al. (1997)

2,4-Dichlorophenoxyacetic acid

Kim et al. (1994, 1995, 1996)

Captan

Fisher et al. (1992); Woollen (1993)

it is known that physiological parameters change with time or a particular exposure scenario, those changes can be easily incorporated into PBPK models. Physicochemical parameters used in PBPK models (i.e., tissue partition coefficients) describe the relative solubility of the chemical in various media (e.g., air, blood, and tissues). Biochemical parameters include the rates of absorption, metabolism, macromolecular binding, and excretion.

To develop a PBPK model, one must first consider which compartments of the organism to include. The com­ partments may be specific organs, anatomical regions, or lumped tissue groups, and their inclusion depends on which animal is being studied and whether or not the com­ partment contributes to the uptake, disposition, and/or tox­ icity of the chemical being modeled. For example, the lung, gastrointestinal tract, and skin may be included because of their ability to serve as sites of absorption, and the kidneys may be included because of their ability to serve as portals of excretion. Tissue partition coefficients and the metabolic capacity of a particular tissue also contribute to chemical disposition. For example, fat is often included in PBPK models due to the high partition coefficient of lipophilic chemicals (e.g., organochlorine pesticides) in adipose tis­ sue, and the liver is often included as a separate compart­ ment because of its involvement in the metabolism of a wide variety of chemicals. Tissues that are target sites for toxicity are often included in PBPK models. The next step in PBPK model development is to write a mass balance equation for each compartment to describe the rate of change of chemical concentration in that com­ partment as a function of time. In the most general case, the mass balance equation for each tissue compartment is rate of change  (rate of uptake)  (rate of removal) The rate of removal is the summation of removal by efflux back into the bloodstream, metabolism within the tissue, and excretion (e.g., biliary excretion in the liver or urinary excretion in the kidney). For simplicity, tissue

Hayes’ Handbook of Pesticide Toxicology

936

compartments that have no capacity for metabolism or excretion will be considered. In this case, the rate of change of chemical concentration in the tissue is the difference in the rates of uptake and efflux. For uptake to occur from blood into a particular tissue compartment, the free chemi­ cal must diffuse out of the capillary space into the intersti­ tial fluid, then diffuse across the plasma membrane to enter the intracellular space (Figure 39.12A). It is assumed that diffusion from the capillary membrane into the interstitial space is very rapid relative to diffusion from the intersti­ tial space into the intracellular space, and the vascular and interstitial subcompartments are represented as one homo­ geneous subcompartment referred to as the extracellular space (Figure 39.12B). Free chemical in the blood enters the extracellular space of the tissue compartment at a rate (mass/time) that is the product of blood flow to the tissue (Qt, units of volume/time) and the concentration of the free chemical in the arterial blood (Ca). Diffusion of the chemi­ cal from the extracellular space across the plasma mem­ brane and into the intracellular space is governed by Fick’s law of diffusion, which states that the rate of transfer of chemical across a membrane (flux) is proportional to its concentration gradient across the membrane flux  PA t C



Vt



dCt  Qt (Ca  Cvt ) dt

(57)

where Vt is the volume of the tissue compartment, Ct is the concentration of free chemical in the tissue compartment, Ca is the concentration of free chemical in the entering arterial blood, and Cvt is the concentration of free chemical in the venous blood exiting the tissue. The concentrations of free chemical in tissue and venous blood leaving the tissue are related by the tissue-to-blood partition coeffi­ cient (Pt) as

(56)

where PAt is the tissue membrane permeation area crossproduct and C is the concentration gradient (units of mass/volume) of free chemical across the membrane.

QtCa

For chemicals that have a perfusion-rate-limited distri­ bution, diffusion of the chemical across the membrane is very rapid relative to its rate of delivery to the tissue (i.e., PAt  Qt), and the rate of uptake by tissues is limited by blood flow rather than the rate of diffusion across the membrane. For such chemicals, the free chemical con­ centration in the intracellular and extracellular spaces is in equilibrium, and the tissue compartment can be repre­ sented as a single homogeneous compartment as shown in Figures 39.11 and 39.12C. The mass balance equation for such a tissue compartment is

Cvt 



Ct Pt

(58)

The overall mixed venous blood concentration is

QtCvt

Vascular space Interstitial space

A

Extracellular space

Intracellular space

QtCa

QtCvt

Extracellular space

B Intracellular space

QtCa

QtCvt

C

Figure 39.12  Uptake from the vascular space of a tissue compartment into the intracellular space where the tissue is represented as three distinct compartments (A), two compartments (B), or a single homogeneous compartment (C). See discussion in Section 39.3.5.

Chapter | 39  Distribution and Pharmacokinetics

Cv 



937

∑ Cvt Qt Qc

(59)

where Qc is cardiac output (i.e., total blood flow or εQt). For chemicals that have a distribution that is limited by the rate of diffusion across the cell membrane rather than blood flow (i.e., diffusion-rate-limited), separate mass bal­ ance equations must be written for the intracellular and extracellular compartments of the tissue (Figure 39.12B). The mass balance equation for the extracellular space is Ves

C  dCes  Qt (Ca  Cvt )  PA t  t  Cvt    Pt dt

(60)

where Ves and Ces are the volume of and the concentration in the extracellular space, respectively. The mass balance equation for the intracellular space (i.e., tissue matrix) is Vis

 dCis C   PA t Cvt  t   dt Pt 

(61)

where Vis and Cis are the volume of and the concentration in the intracellular space, respectively. Chemicals may be effectively eliminated from a tissue by metabolism, macromolecular binding, and/or excre­ tion. For these tissues, the mass balance equation is more complex than those shown in Eqs. (57), (60), and (61). The mass balance equation for a tissue that metabolizes the chemical (e.g., the liver) is Vt



dCt dA  Qt (Ca  Cvt )  met dt dt

(62)

where dAmet/dt is the rate of metabolism. Many enzyme systems are saturable, and d Amet V C  max vt dt K m  Cvt



(63)

where Vmax and Km are the maximal velocity and Michaelis constant of the enzymatic reaction. First-order metabolism is described by the equation dAmet  Vt K f Cvt dt





V C dCl  Ql (Ca  Cvl )  max vl  K f CvlVl dt K m  Cvl

Ca 

Qc Cv  QpCi Qc  (Qp / Pb )



(65)

(66)

where Qp is alveolar ventilation rate, Ci is the concentration in inhaled air, and Pb is the blood:air partition coefficient. If the chemical can be inhaled, it can also be exhaled. The concentration of the chemical in exhaled alveolar air is the ratio of its concentration in arterial blood and the blood:air partition coefficient, Ca/Pb. Upon oral ingestion of a chemical (e.g., food, drinking water), the chemical may be absorbed into the portal cir­ culation or into the lymphatic system. For simplicity, we assume first-order absorption into the portal circulation only. In this situation, the chemical is delivered directly to the liver prior to distribution throughout the body. This requires an input term in the mass balance equation for the liver and a mass balance equation that describes the rate of loss of chemical from the site of absorption (stomach): dAst  K a Ast dt





Vl

(67)

V C dCl  Ql (Ca  Cvl )  Ka Ast  max vt  K f Cvl Vl dt K m  Cvt

(68)

where Ka is the oral absorption rate constant and Ast is the amount (mass) of chemical in the stomach (i.e., the site of absorption). Incorporation of dermal absorption is more complex in that a skin compartment must be added to the model (see Krishnan and Andersen, 1994). In addition to metabolism, chemicals may be eliminated from the body by excretion in urine, exhaled air, and other routes that are not be discussed here (e.g., sweat, bile, and milk). A typical equation for the concentration of chemical in exhaled air is

(64)

where Kf is the first-order rate constant. Both saturable and first-order metabolism may occur simultaneously. For example, if metabolism by both saturable and first-order processes is occurring in the liver, the mass balance equa­ tion for the liver would be Vl

For inhalation exposure to volatile chemicals, equilib­ rium is established between the chemical in the alveolar air space of the lung and the chemical in arterial blood. The concentration of chemical in arterial blood (Ca) is described by the equation



Cex  0.7

Ca  0.3Ci Pb

(69)

Equation (69) uses the assumption that exhaled air is a mixture of inhaled air (30%) and expired alveolar air (70%). Urinary excretion may be described in a number of ways, including excretion by first-order and saturable processes similar to the descriptions of metabolism in Eqs. (63) and (64). The same is true for elimination by other routes.

938

Conclusion In order to either exert a deleterious effect or be detoxi­ fied, xenobiotics, including pesticides, must be absorbed, distributed, metabolized, react with a macromolecule or other receptor, and/or be excreted. Distribution and phar­ macokinetics are the subjects of this chapter. In the case of toxicants such as pesticides, the term toxicokinetics is frequently used. Noncompartmental, compartmental, and physiological models are considered, including the quanti­ tative methodology for utilizing these models. The mecha­ nisms of absorption and transport, metabolism, mode of toxic action, and excretion are considered elsewhere in this handbook.

References Ambrose, A. M., Christensen, H. E., Robbins, D. J., and Rather, L. J. (1953). Toxicological and pharmacological studies on chlordane. Arch. Ind. Hyg. Occup. Med. 7, 197–210. Anadon, A., Martinez-Larranaga, M. R., Diaz, M. J., and Bringas, P. (1991). Toxicokinetics of permethrin in the rat. Toxicol. Appl. Pharmacol. 110, 1–8. Arms, A. D., and Travis, C. C. (1988). “Reference Physiological Parameters in Pharmacokinetic Modeling.” Rep. NTIS PB 88-196019, Office of Health and Environmental Assessment, U.S. Environmental Protection Agency, Washington, DC. Burgaz, S., Afkham, B. L., and Karakaya, A. E. (1994). Organochlorine pesticide contaminants in human adipose tissue collected in Ankara (Turkey) 1991–1992. Bull. Environ. Contam. Toxicol. 53, 501–508. Casey, P. H., and Collie, W. R. (1984). Severe mental retardation and mul­ tiple congenital anomalies of uncertain cause after extreme parental exposure to 2,4-D. J. Pediatr. 104, 313–315. Chui, Y. C., Poon, G., and Law, F. (1988). Toxicokinetics and bioavail­ ability of paraquat in rats following different routes of administration. Toxicol. Ind. Health 4, 203–219. Dale, W. E., Gaines, T. B., and Hayes, W. J. Jr. (1962). Storage and excre­ tion of DDT in starved rats. Toxicol. Appl. Pharmacol. 4, 89–106. Davidow, B., and Frawley, J. P. (1951). Tissue distribution, accumulation and elimination of the isomers of benzene hexachloride. Proc. Soc. Exp. Biol. Med. 76, 780–783. DeJongh, J., and Blaauboer, B. J. (1997). Simulation of lindane kinetics in rats. Toxicology 122, 1–9. Durham, W. F. (1969). Body burden of pesticides in man. Ann. NY Acad. Sci. 160, 183–195. el-Masri, H. A., Thomas, R. S., Benjamin, S. A., and Yang, R. S. H. (1995). Physiologically based pharmacokinetic/pharmacodynamic modeling of chemical mixtures and possible applications in risk assessment. Toxicology 105, 275–282. el-Masri, H. A., Thomas, R. S., Sabados, G. R., Phillips, J. K., Constan, A. A., Benjamin, S. A., Andersen, M. E., Mehendale, H. M., and Yang, R. S. H. (1996). Physiologically based pharmacokinetic/phar­ macodynamic modeling of the toxicologic interaction between car­ bon tetrachloride and Kepone. Arch. Toxicol. 70, 704–713. Fisher, H. L., Hall, L. L., Sumler, M. R., and Shah, P. V. (1992). Dermal penetration of [14C]captan in young and adult rats. J. Toxicol. Environ. Health 36, 251–271.

Hayes’ Handbook of Pesticide Toxicology

Freeman, R. A., Rozman, K. K., and Wilson, A. G. E. (1989). Physiological pharmacokinetic model of hexachlorobenzene in the rat. Health Phys 57(Suppl. 1), 139–147. Garcia-Repetto, R., Martinez, D., and Repetto, M. (1995). Coefficient of distribution of some organophosphorous pesticides in rat tissue. Vet. Human Toxicol. 37, 226–229. Gearhart, J. M., Jepson, G. W., Clewell, H. J., Andersen, M. E., and Conolly, R. B. (1994). Physiologically based pharmacokinetic model for the inhibition of acetylcholinesterase by organophosphate esters. Environ. Health Perspect. 102(Suppl. 11), 51–60. Gibaldi, M., and Perrier, D. (1982). “Pharmacokinetics,” 2nd ed. Dekker, New York. Hayes, W. J. Jr., Quinby, G. E., Walker, K. C., Elliott, J. W., and Upholt, W. M. (1958). Storage of DDT and DDE in people with different degrees of exposure to DDT. Arch. Ind. Health 18, 398–406. Hissink, A. M., Van Ommen, B., Kruse, J., and Van Bladeren, P. J. (1997). A physiological based pharmacokinetic (PB-PK) model for 1,2-dichlorobenzene linked to two possible parameters of toxicology. Toxicol. Appl. Pharmacol. 145, 301–310. Kim, C. S., Binienda, Z., and Sandberg, J. A. (1996). Construction of a physiologically based pharmacokinetic model for 2,4-dichlorophen­ oxyacetic acid dosimetry in the developing rabbit brain. Toxicol. Appl. Pharmacol. 136, 250–259. Kim, C. S., Gargas, M. L., and Andersen, M. E. (1994). Pharmacokinetic modeling of 2,4-dichlorophenoxyacetic acid (2,4-D) in rat and in rabbit brain following single dose administration. Toxicol. Lett. 74, 189–201. Kim, C. S., Slikker, W. Jr., Binienda, Z., Gargas, M. L., and Andersen, M. E. (1995). Development of a physiologically based pharmacoki­ netic model for 2,4-dichlorophenoxyacetic acid dosimetry in discrete areas of the rabbit brain. Neurotoxicol. Teratol. 17, 111–120. Krishnan, K., and Andersen, M. E. (1994). Physiologically based pharmaco­ kinetic modeling in toxicology. In “Principles and Methods of Toxicology” (A. W. Hayes, ed.), 3rd ed., pp. 149–188. Raven Press, New York. Lehman, A. J. (1956). The minute residue problem. Q. Bull. Assoc. Food Drug Off. 20, 95–99. Leung, H. W., and Paustenbach, D. J. (1988). Application of pharmaco­ kinetics to derive biological exposure indexes from threshold limit values. Am. Ind. Hyg. Assoc. J. 49, 445–450. Ludwig, G., Weis, J., and Korte, F. (1964). Excretion and distribution of aldrin-14C and its metabolites after oral administration for a long period of time. Life Sci. 3, 123–130. Robinson, J., Roberts, M., Baldwin, M., and Walker, A. I. T. (1969). The pharmacokinetics of HEOD (dieldrin) in the rat. Food Cosmet. Toxicol. 7, 317–332. Roth, W. L., Freeman, R. A., and Wilson, A. G. E. (1993). A physiologi­ cally based model for gastrointestinal absorption and excretion of chemicals carried by lipids. Risk Anal. 13, 531–543. Sandberg, J. A., Duhart, H. M., Lipe, G., Binienda, Z., Slikker, W. Jr., and Kim, C. S. (1996). Distribution of 2,4-dichlorophenoxyacetic acid (2,4-D) in maternal and fetal rabbits. J. Toxicol. Environ. Health 49, 497–509. Street, J. C. (1964). DDT antagonism to dieldrin storage in adipose tissue of rats. Science 146, 1580–1581. Thies, M. L., and McBee, K. (1994). Cross-placental transfer of organo­ chlorine pesticides in Mexican free-tailed bats from Oklahoma and New Mexico. Arch. Environ. Contam. Toxicol. 27, 239–242. Timchalk, C., Dryzga, M. D., Johnson, K. A., Eddy, S. L., Freshour, N. L., Kropscott, B. E., and Nolan, R. J. (1996). Comparative phar­ macokinetics of [14C]metosulam (N-[2,6-dichloro-3-methylphenyl]-5, 7-dimethoxy-1,2,4-triazolo[1,5a]-pyrimidine-2-sulfonamide) in rats, mice and dogs. J. Appl. Toxicol. 17, 9–21.

Chapter | 39  Distribution and Pharmacokinetics

Wagner, J. G. (1967). Drug accumulation. J. Clin. Pharmacol. 7, 84–88. Wagner, J. G., Northam, J. I., Alway, C. D., and Carpenter, O. S. (1965). Blood levels of drug at the equilibrium state after multiple dosing. Nature 207, 1301–1302. Woollen, B. H. (1993). Biological monitoring for pesticide absorption. Ann. Occup. Hyg. 37, 525–540. Wu, H. X., Evreux-Gros, C., and Descotes, J. (1996). Diazinon toxico­ kinetics, tissue distribution and anticholinesterase activity in the rat. Biomed. Environ. Sci. 9, 359–369.

939

Yang, R. S., el-Masri, H. A., Thomas, R. S., and Constan, A. A. (1995a). The use of physiologically-based pharmacokinetic/pharmacodynamic dosimetry models for chemical mixtures. Toxicol. Lett. 82–83, 497–504. Yang, R. S., el-Masri, H. A., Thomas, R. S., Constan, A. A., and Tessari, J. D. (1995b). The application of physiologically based pharmacokinetic/ pharmacodynamic (PBPK/PD) modeling for exploring risk assessment approaches of chemical mixtures. Toxicol. Lett. 79, 193–200. Zatz, J. L. (1972). Accumulation of organochlorine pesticides in man. J. Pharm. Sci. 61, 948–949.

Chapter 40

Metabolic Interactions of Pesticides Ernest Hodgson North Carolina State University, Raleigh, North Carolina

40.1  Chemical factors affecting pesticide metabolism: introduction Although the study of the metabolism and toxicity of pesticides is simplified by considering single compounds, humans and other living organisms are not exposed in this way; rather, they are exposed to many xenobiotics simultaneously, involving different portals of entry, modes of action, and metabolic pathways. Because they bear directly on the problem of toxicity-related interactions between different xenobiotics, metabolic interactions between exogenous compounds are important in the study of pesticide toxicity. Also of importance in considerations of pesticide toxicity and safety are metabolic interactions between pesticides and endogenous metabolites. Pesticides, and other xenobiotics, in addition to serving as substrates for a number of xenobiotic-metabolizing enzymes (XMEs), may also serve as inhibitors or inducers of these or other enzymes. Moreover, there are many compounds that first inhibit and subsequently induce such enzymes as the microsomal monooxygenases. The situation is even further complicated by the fact that, although some substances have an inherent toxicity and are detoxified in the body, others without inherent toxicity can be metabolically activated to potent toxicants. The following examples are illustrative of the situations that might occur involving two compounds: Compound A, without inherent toxicity, is metabolized to a potent toxicant. In the presence of an inhibitor of its metabolism, there would be a reduction in toxic effect, while after exposure to an inducer of the activating enzymes, there would be an increase. Conversely, the toxicity of compound B, a toxicant that is metabolically detoxified, would be increased in the presence of an inhibitor and decreased in the presence of an inducer. In addition to these possible cases, the toxicity of the inhibitor or inducer, as well as the time dependence of the effect, must also be considered, because, as mentioned previously, many xenobiotics that are initially enzyme inhibitors ultimately become inducers. Interactions between Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

components in mixtures more complex than binary mixtures represent a particularly intractable program that has not yet been well resolved. Other xenobiotics, such as clinical or other drugs or occupational chemicals, by causing enzyme induction or inhibition, can affect the metabolism and thus the toxicity of pesticides. Conversely, pesticides, by acting as either enzyme inducers or inhibitors, can affect the metabolism of other xenobiotics, such as drugs, as well as the metabol­ ism of endogenous compounds, such as steroid hormones. In the following sections, the discussion and examples will serve to illustrate these various interactions. Although the mechanisms of enzyme inhibition and induction are investigated by a variety of biochemical and molecular biological techniques, it is important, for consideration of the implications of these phenomena in human health risk assessment, to demonstrate them in vivo. Some examples of methods and chemicals used for this purpose are shown in Table 40.1.

40.2  Induction 40.2.1  Induction of Microsomal Enzyme Activity The stimulatory effect of xenobiotics on liver microsomal enzymes was first reported in the 1950s (Brown et al., 1954; Conney et al., 1957; Miller et al., 1954; Remmer, 1958) and since then has been extensively investigated. Numerous early experiments with laboratory rodents confirmed hepatic enzyme induction, although until recently methods were not available for identification of individual isoforms and the inducer was often classified as a phenobarbital-, a 3-methyl-cholanthrene-, or a mixed-type inducer. Reviews in this area include those of Conney (1967), Gelboin and Conney (1968), Sher (1971), Gillette et al. (1972), Nebert and Jensen (1979), Okey et al. (1986), Okey (1990), Batt et al. (1992), and Denison and Whitlock 941

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942

Table 40.1  In Vivo Assessment of Altered Microsomal Activities in Humans and Animals Using Test Compounds Test compound

Test

Species

Reference

Aminopyrene

Breath test

Human

Jager et al. (1980)

Antipyrine

Plasma half-life, urinary excretion

Human Rat

Mehta et al. (1982) Butler and Dauterman (1989)

Caffeine

Plasma half-life, urinary excretion

Human

Kadlubar et al. (1992) Relling et al. (1992)

Chloramphenicol

Plasma half-life

Human

Mehta et al. (1975)

Hexobarbital

Sleep time

Rat

Butler and Dauterman (1988)

-Methyldigoxin

Plasma half-life, urinary excretion

Human

Hinderling and Garrett (1977)

Phenylbutazone

Plasma half-life

Human

Krishnaswamy et al. (1981)

Procaine

Paralysis

Rat

Butler and Dauterman (1988)

Salicylates

Plasma half-life, urinary excretion

Rat

Yu and Varma (1982)

Theophylline

Plasma half-life

Human Rat

Mehta et al. (1982) Butler and Dauterman (1988)

(1995). Reviews with emphasis on pesticides include those of Fouts (1963), Conney et al. (1967), Leibman (1968), Street et al. (1969), DuBois (1969), Hodgson (1974), Hodgson and Kulkarni (1974), Hodgson et al. (1980), Wilkinson and Denison (1982), Khan (1984), Kulkarni and Hodgson (1984a,b), Hodgson and Levi (1996), and Hodgson and Meyer (1997, 2009). It should be noted that induction is not restricted to xenobiotics, and enzymes may also be induced by hormones and other normal body constituents (Conney, 1967; Conney et al., 1967, 1979; Kobliakov et al., 1991; Pantuck et al., 1979, 1984; Ronis and Cunny, 1994; Schenkman et al., 1989) and by dietary constituents (Anderson and Kappas, 1991; Donaldson, 1994; Hodgson and Meyer, 2009; Wattenberg, 1971). More recently, human hepatocytes have been used as a model system for investigating induction of microsomal monooxygenases and other XMEs by pesticides (e.g., Das et al., 2006, 2008a,b). A large number of studies have provided evidence for induction of enzymes in surrogate animals or in humans who have been exposed occupationally or environmentally to pesticides (Table 40.2). For the most part these studies employed noninvasive in vivo techniques such as examination of the half-life of aminopyrene or phenylbutazone or excretion of 6-hydroxycortisol (Guzelian et al., 1980; Kolmodin et al., 1969; Kolmodin-Hedman, 1973; Kreiss et al., 1981; Poland et al., 1970). More recently, in vitro techniques are used following in vivo exposure, techniques that yield information on such aspects as isoform specifi­ city and the mechanism of induction. For example, Dalton et al. (2003), Das et al. (2006, 2008a,b), and Johri et al. (2006, 2008) are among the many examples cited throughout this chapter.

Abernathy et al. (1971a,b) demonstrated significant decreases in zoxazolamine paralysis time, hexobarbital sleeping time, and aniline hydroxylase activity in mice following treatment with dichlorodiphenyltrichloroethane (DDT) or 1,1-dichloro-2,2-bis(p-chlorophenyl)ethylene (DDE), a major metabolite of DDT and a persistent residue in animals, including humans, even in countries where DDT use has been banned for decades. Different inducers may increase the expression of different enzymes and, therefore, different metabolic pathways. Thus, Chadwick et al. (1971) showed that repeated doses of lindane or DDT increased oxidative hydrolysis, O-demethylase, dehydrochlorinase, and glucuronyl transferase activity, but to different degrees. Pretreatment of rats with lindane caused them to metabolize a single dose of radioactive lindane 2.5 times more extensively than controls, and pretreatment with DDT caused a 3.5-fold increase in metabolism of radioactive lindane. Furthermore, the DDT pretreatment was followed by proportionally more neutral and weakly polar, but less free-acid-type metabolites of the radioactive lindane. Thus, metabolism was qualitatively as well as quantitatively different following administration of the two inducers. Subsequent studies (Chadwick and Freal, 1972) confirmed these findings, including the increased excretion of metabolites following pretreatment with DDT. In addition, it was shown that rats pretreated with DDT plus lindane excreted more 2,4,5-trichlorophenol and 2,3,4,6- and 2,3,4,5-tetrachloro­ phenols by the second day of treatment than did rats receiving lindane alone. The results suggested that DDT treatment stimulates the metabolism of lindane through a selective effect on certain metabolic pathways involved in the oxidative degradation of lindane, notably those leading

Inducer and/or substrate

Species

Effect

Reference

Mouse Mouse

Hexobarbital sleeping time increased Hexobarbital sleeping time increased Hexobarbital sleeping time increased up to12 hours, decreased after 24–72 hours Parathion toxicity increased after 1 hour, decreased after 48 hours Microsomal CYP content decreased after 2–12 hours, increased after 12–36 hours Induction of CYP 2B10, 1A1, and 1A2; 1A2 by an Ah-independent mechanism

Fine and Molloy (1964) Fine and Molloy (1964) Kamienski and Murphy (1971)

Synergists Sesoxane Piperonyl butoxide

Kamienski and Murphy (1971) Philpot and Hodgson (1972) Philpot and Hodgson (1972); Lewandowski et al. (1990); Adams et al. (1993a,b, 1995); Ryu et al. (1996); Cook and Hodgson (1985, 1986)

Chlorinated hydrocarbon insecticides BHC

Rat

In vitro metabolism of hexobarbital increased, hexobarbital sleeping time decreased, scillicocide toxicity decreased. All isomers similar

Koransky et al. (1964)

Trichloro-237

Rat

Hexobarbital sleeping time decreased

Hart and Fouts (1963)

-Chlordane

Rat

Hexobarbital sleeping time decreased

Hart and Fouts (1963)

Endrin

Rat

Hexobarbital sleeping time decreased

Hart and Fouts (1963)

DDT

Rat

No effect on hexobarbital sleep time Hexobarbital sleep time decreased. Hexobarbital metabolism in vitro increased. Metabolism of aminopyrine increased. Metabolism of p-nitrobenzoic acid increased. No effect on aniline metabolism. Increased detoxication of EPN. Increased O-demethylation of p-nitroanisole. Increased N-demethylation of aminopyrine Increased metabolism, in vitro, of EPN and p-nitroanisole Increase in CNS arousal with phenobarbital administration

Hart and Fouts (1963) Hart and Fouts (1963); Kinoshita et al. (1966)

Squirrel monkey Human

Chapter | 40  Metabolic Interactions of Pesticides

Table 40.2  Induction of Microsomal Enzyme Activity Following Treatment, in Vivo, and Involving Pesticides as Either Inducers or Substrates

Cranmer et al. (1972) Rappolt (1973)

DDT and analogues

Mouse

Increased P450 levels, aniline hydroxylase activity, zoxazolamine paralysis time and hexobarbital sleep time. QSAR for 28 analogues

Abernathy et al. (1971a,b)

Diphenyl hydantoin

Rat

DDT and DDE storage decreased, in vitro DDT metabolism increased

Cranmer (1970)

o,p-DDD

Guinea pig

Phenobarbital sleep time decreased, in vitro phenobarbital metabolism increased

Straw et al. (1965) (Continued)

943

944

Table 40.2  (Continued) Inducer and/or substrate

Species

Effect

Reference

p,p-DDD o,p-DDT

Rat Rat HepG2 cells

Increased metabolism in vitro of estradiol-17 PXR- and CAR-dependent increase in CYPs 2B2 and 3A2 Increase in CYP3A4 mRNA

Welch et al. (1971) Kiyosawa et al. (2008) Medina-Diaz et al. (2005, 2007)

p,p-DDE

Rat

Increased metabolism in vitro of estradiol-17 Induction of CYP 2b 3A

Welch et al. (1971) Wyde et al. (2003)

Chlordane

Rat

In vitro metabolism of hexobarbital, aminopyrine, and chlorpromazine unchanged after one dose; all increased after three doses Decreased toxicity of dicoumarol Increased metabolism in vitro of estrone

Hart and Fouts (1963)

Dog

Welch and Harrison (1966) Welch et al. (1971)

Rat

Increased metabolism in vitro of estradiol-17 Increased expression of mRNA for CYP1A1, 1A2, 2B1, 2B2, and 2E1 as well as associated catalytic activities

Welch et al. (1971) Johri et al. (2003, 2008)

Heptachlor

Rat

Increased metabolism in vitro of estradiol-17

Welch et al. (1971)

Toxaphene

Rat

Increased metabolism in vitro of estradiol-17 Increased detoxication of EPN, O-demethylation of p-nitroanisole and N-demethylation of aminopyrine

Welch et al. (1971) Kinoshita et al. (1966)

Dieldrin

Rhesus monkey Dog Rat Rat Mouse

Increased metabolism in vitro of chlorfenvinphos Increased metabolism in vitro of chlorfenvinphos Increased metabolism in vitro of chlorfenvinphos Increased metabolism in vitro of estradiol-17 Increased metabolism in vitro of chlorfenvinphos

Wright et al. (1972) Wright et al. (1972) Wright et al. (1972) Welch et al. (1971) Wright et al. (1972)

Mirex

Mouse/rat

Increased O-demethylation in vitro of p-nitroanisole and CYP content

Baker et al. (1972)

Mirex and kepone

Rat

Increased warfarin hydroxylation in vitro and increased CYP content Increased benzo(a)pyrene hydroxylase activity in vitro and increased CYP content Increased acute in vivo hepatotoxicity; mirex had greater effect Increased CYP and N- and O-dealkylation in vitro Induction of CYP2B10, 1A2, and 3A

Kaminisky et al. (1978)

Increase in total CYP and gender-dependent increases in CYP 1A1- and 1A2-related activities

Orepeza-Hernandez et al. (2003)

Gerbil Mouse

Methoxychlor

Rat

Crouch and Ebel (1987) Fouse and Hodgson (1987) Fabacher and Hodgson (1976) Lewandowski et al. (1989)

Hayes’ Handbook of Pesticide Toxicology

Lindane

Chapter | 40  Metabolic Interactions of Pesticides

Organophosphorus insecticides 3-Methylcholanthrene

Rat

Increased metabolism in vitro of azinphosmethyl to a cholinesterase inhibitor

Murphy and DuBois (1957)

Phenobarbital

Rat/mouse

Decreased in vivo toxicity of parathion, methyl parathion, demeton, disulfoton, azinphosmethyl, dioxathion, ethion, carbophenothion, mevinphos, and EPN

Dubois (1969)

Malathion

Rat

Increased CYP

Matthews and Devi (1994)

S,S,S-Tri-n-butyl phosphorotrithioate

Hen

Increased CYP

Lapadula et al. (1984)

Chlorpyrifos

Human

Increased CYP isoforms in hepatocytes

Das et al. (2008)

Chicken

Pentobarbital sleep time decreased

Puryear and Paulson (1972)

Permethrin

Rat, human

Increased CYP isoform protein and mRNA in hepatocytes

Heder et al. (2001); Das et al. (2008)

Deltamethrin

Rat, human

Increase in CYP mRNA and protein

Johri et al. (2006); Das et al. (2008)

Pyrethrins

Rat

Increase in CYP2B1 and CYP2B1/2 mRNA and associated enzymatic activities; also increase in testosterone 6-hydroxylase activity Increase in testosterone 6-hydroxylase activity, CYP2B6 and CYP3A4 mRNA

Price et al. (2008)

Carbamate insecticides Carbaryl Pyrethroid insecticides

Human

Price et al. (2008)

Rodenticides Phenobarbital

Human Dog

Pharmacological activity of warfarin decreased Dicoumarol toxicity in vivo decreased

Robinson and MacDonald (1966) Welch et al. (1967)

Acetylsalicylic acid

Rat

Dicoumarol prothrombin time decreased

Coldwell and Zawidzka (1968)

Heptobarbital

Human

Excretion of dicoumarol metabolites increased

Aggeler and O’Reilly (1969)

Alachlor

Rat

Induction of CYP2B1/2 and 1A1/2 protein and associated activities

Hanioka et al. (2002)

Metolachlor

Rat

Induction of CYP2B1/2 and CYP3A1/2 protein

Dalton et al. (2003)

Herbicides

(Continued)

945

946

Table 40.2  (Continued) Inducer and/or substrate

Species

Effect

Reference

Monuron 1967

Rat

Detoxication of EPN, O-demethylation of p-nitroanisole and N-demethylation of aminopyrine increase for 1-3 weeks, then return to normal

Kinoshita and DuBois (1967)

Diuron

Rat

Detoxication of EPN, O-demethylation of p-nitroanisole and N-demethylation of aminopyrine Mouse hepatoma cells increase for 1-3 weeks, then return to normal AhR-dependent induction of CYP1A1 mRNA

Kinoshita and DuBois (1967) Zhao et al. (2006)

Tridiphane

Mouse

Induction of CYP4A Induction of epoxide hydrolase

Levi et al. (1992) Moody and Hammock (1987)

Griseofulvin

Human

Pharmacological action of warfarin decreased

Cullen and Catalano (1967)

Parnon

Rat

In vitro metabolism increased

Hoffman et al. (1968)

Azoles

Rat

Increase in liver weight, liver pathology, total CYP, CYPs 2B1 and 3A2, and associated enzyme activities Increase in liver weight, liver pathology, total CYP, CYPs 2B1 and 3A2, and associated enzyme activities

Barton et al. (2006); Sun et al. (2006, 2007); Martin et al. (2007) Allen et al. (2006); Ward et al. (2006); Sun et al. (2006)

Human

Increased CYP isoforms in hepatocytes

Das et al. (2006)

Mice

PPAR-dependent increases in peroxisomal acyl-CoA oxidase, thiolase, and Cyp4a10 and4a14

Upham et al. (2007)

Human

Increased CYP isoforms in hepatocytes

Das et al. (2008)

Fungicides

Mice Pyrazole insecticide Fipronil Pesticide adjuvant Toximul

DEET

Adapted from Hodgson and Meyer (2009).

Hayes’ Handbook of Pesticide Toxicology

Insect repellent

Chapter | 40  Metabolic Interactions of Pesticides

to the formation of tetrachlorophenols, particularly 2,3, 4,5-tetrachlorophenol. Incidentally, when two inducers are involved, the resulting induction may be either additive or slightly anta­ gonistic. Thus, Gielen and Nebert (1971) found an additive effect when either phenobarbital or p,p-DDT was present with a polycyclic hydrocarbon, but not when combinations of phenobarbital plus DDT or one polycyclic hydrocarbon plus another were involved. Although it had been known for many years that various pesticides could induce cytochrome P450s (CYPs), neither the specific isozymes induced nor the implications of the induction were well characterized. Later studies using enzymatic and immunochemical techniques examined isoform specificity of specific pesticides. For example, mirex and chlordecone were shown to induce CYP2B10 and testosterone metabolism in mouse liver, a pattern of induction similar to that of phenobarbital (Baker et al., 1972; Fabacher and Hodgson, 1976; Hodgson, 1974; Lewandowski et al., 1989). Enzymatic activities suggested that, in addition to 2B10, other CYPs were induced, and later studies demonstrated induction of CYP1A2 and CYP3A (Dai et al., 1998; Hodgson and Levi, 1996). Another group of pesticides, the phenoxyacetic acid herbicides (e.g., 2,4-dichlorophenoxyacetic acid [2,4-D]), and the herbicide synergist tridiphane were found to induce the CYP4A isozymes in rodents (Levi et al., 1992; Moody et al., 1991, 1992). These CYP isoforms are known to be involved in the oxidation of fatty acids and the maintenance of lipid homeostasis. Moreover, in rodents, compounds that are CYP4A inducers also cause peroxisome proliferation, an event associated with nongenotoxic induction of liver tumors in rodents. Another peroxisome proliferator, the fenvalerate metabolite fenvaleric acid, has been shown to induce several CYP-dependent enzyme activities including 7-ethoxyresorufin-deethylation, catalyzed by CYP1As, 7-pentoxyresorufin O-dealkylation, catalyzed by CYP2Bs, and testosterone hydroxylation, catalyzed by CYP3A and CYP2B11 (Morisseau et al., 1991). The exact relationship of these interactions and the relevance to humans has not yet been defined. Methylenedioxyphenyl (MDP) compounds, such as piperonyl butoxide (PBO) and sesamex (SES), have been used as synergists with pyrethroid and carbamate pesticides. Other well-known MDP compounds such as safrole and isosafrole are found in many common foods of plant origin. These chemicals affect multiple enzyme pathways, including the CYP system (Hodgson and Philpot, 1974). Their effect on CYP enzymes is biphasic, that is, inhibition followed by induction, and is discussed more fully in Section 40.4. Reviews include those of Philpot and Hodgson (1971–1972), Hodgson and Levi (1998), and Hodgson (1999). However, MDP compounds are known to induce both enzyme and mRNA for several CYP isoforms in the mouse, including CYP1A1, CYP1A2, and

947

CYP2B10. CYP1A2 is induced by both Ah-receptordependent and Ah-receptor-independent mechanisms (Cook and Hodgson, 1985, 1986; Ryu et al., 1995, 1996). The fungicide captan, although inhibiting many hepatic CYP-dependent activities in mouse liver (Paolini et al., 1999), induces both CYP3A and CYP1A2 in the kidney and CYP1A2 in the lung. The ergosterol biosynthesis inhibiting fungicides (EBIFs), for example, clotrimazole and propioconazole, have been shown to have multiple effects on the rodent CYP system (Ronis et al., 1994). The EBIFs induced CYPs 3A, 2B, and 1A, while suppressing the activity of CYP2C11. These alterations were found to cause significant changes in testosterone metabolism in male rats. Cellular techniques are becoming more available for the study of induction by pesticides. The azole fungicides have been shown to be inducers of various XMEs, primarily in rodents (Allen et al., 2006; Barton et al., 2006; Martin et al., 2007; Sun et al., 2006, 2007; Ward et al., 2006). Among recent studies of interest is the demonstration of the induction of CYP4A10 and CYP4A14 in mice by the pesticide adjuvant, Toximul, an effect mediated through the PPAR receptor (Upham et al., 2007). Diuron and related phenylurea herbicides induced CYP1A1 via the Ah receptor in several cell lines, including at least one human cell line (Zhao et al., 2006). Further studies showed the induction, in rodents, of CYPs 2B and 3A by the herbicide metolachlor (Dalton et al., 2003), of CYPs 1A and 2B by deltamethrin (Johri et al., 2006), and of a number of CYP-related metabolic activities by the herbicide alachlor (Hanioka et al., 2002). DuBois et al. (1996) used hepatocytes from rat and quail as well as human hepatoma (HepG2) cells to study induction of CYP isoforms by pesticides. The pesticides fell into four groups: first, CYP3A inducers such as pentachlorophenol; second, 3-methylcholanthrene-type inducers, such as lindane, an inducer of CYP1A isoforms; third, phenobarbitaltype inducers, such as dieldrin, an inducer of CYP2B isoforms; and fourth, pesticides with little or no capacity to induce CYP isoforms. Pentachlorophenol and lindane were the strongest inducers in these cell lines, and lindane appeared to be a member of both the second and third groups because it induced both CYP1A and CYP2B activities. Although a small number of studies showed induction by pesticides in rat hepatocytes, for example, the induction of CYP2B1 by pyrethroids (Heder et al., 2001), the recent availability of human hepatocytes has enabled the investigation of induction by pesticides in humans (Das et al., 2006, 2008a,b). These studies reveal that fipronil is an effective inducer of CYP1A1, CYP2B6, and CYP3A4 in human hepatocytes, at concentrations as low as 1.0 M. The pyrethroids, deltamethrin amd permethrin, while not as effective as fipronil, induced the same CYP isoforms. The insecticide chlorpyrifos and the insect repellent DEET are also capable of inducing xenobiotic-metabolizing CYP

948

isoforms in human hepatocytes. All of these results of studies utilizing human hepatocytes suggest that pesticide– pesticide interactions are possible in the human liver, and this possibility should be investigated. It might also be noted that, to a greater or lesser extent, these pesticides are also cytotoxic to human hepatocytes but generally at higher concentrations than those required for CYP isoform induction. Of particular concern is the ability of many pesticides to disrupt the normal functioning of the endocrine system, and endocrine disruption has become an important environmental concern (Birnbaum, 1994; Colborn et al., 1993; Guillette et al., 1994). For this reason, many of the recent studies of induction by organochlorines have involved either methoxychlor or o,p-DDT. Methoxychlor has been shown to induce CYPs 1A, 2B, 2C, 2E, and 3A in both male and female rats (Oropeza-Hernandez, 2003) and to induce ethoxyresorufin O-deethylase (EROD) and pentoxyresorufin O-depentylase (PROD) activity in HepG2 cells (Dehn et al., 2005; MedinaDiaz and Elizondo, 2005). o,p-DDT or its metabolite o,pDDE have been shown to induce CYP3A4 (Medina-Diaz et al., 2007). The finding that the nuclear receptors pregnane X receptor (PXR) and constitutive androstane receptor (CAR) are involved in induction in both HepG2 cells and in immature ovariectomized rats (Kiyosawa et al., 2008; Wyde et al., 2003) is of mechanistic importance. Lindane has been shown to induce CYPs 1A and 2B in rats (Johri et al., 2008; Parmar et al., 2003). It is now quite clear that pesticides may have complex effects on the hepatic monooxygenase system and, in addition to affecting xenobiotic metabolism, pesticides, by disturbing endogenous metabolism, have the potential to result in profound changes in both the physiological and reproductive capacities of the organism (Birnbaum, 1994; Tyler et al., 1998).

40.2.2  Induction of Other Enzymes Microsomal enzymes are not the only ones subject to induction. For example, -aminolevulinic acid (ALA) synthetase is located in the mitochondria and may increase 40– 100 times in those structures on induction (Granick, 1965). Interaction between induction of mitochondrial and microsomal enzymes is illustrated by the action of the pesticide m-dichlorobenzene in rats. Following daily doses at the rather high rate of 800 mg/kg, there is a biphasic stimulation of ALA-synthetase activity and of the excretion of urinary coproporphyrin, both of which peak by 3 days and then decline. The decrease in ALA-synthetase and in excretion of coproporphyrin at 5 days corresponds with the maximal stimulation of drug metabolism and with a decrease in the concentration of m-dichlorobenzene in the serum and liver at the time (Poland et al., 1971). In the cytosolic fraction of homogenized rat liver, the activity of nicotinamide adenine dinucleotide (NAD)-dependent aldehyde dehydrogenase (EC 1.2.1.3) is increased up to

Hayes’ Handbook of Pesticide Toxicology

10-fold after administration of phenobarbital for 3 days. The effect is genetically controlled and is inherited as an autosomal dominant characteristic. The mechanism is apparently unrelated to other drug-induced increases in enzyme activity such as those that occur in the hepatic microsomal systems for drug metabolism (Deitrich, 1971). Glutathione S-transferases as well as CYP were shown to be induced by pesticides, but the levels of induction of the former were much lower (Fabacher et al., 1980; Hodgson et al., 1980; Kulkarni et al., 1980; Robacker et al., 1981).

40.2.3  Mechanism of Induction Several pesticides have been tested for activation of the PXR receptor in engineered expression systems (Lemaire et al., 2006; Matsubara et al., 2007). Since human 3A4 is transcriptionally regulated, in part, by PXR, induction of this CYP is expected for the pesticides active in the in vitro systems. In rats, one of these compounds, the herbicide metolachlor, was shown to induce the rat ortholog CYP3A2 as well as CYPB1/2 at approximately one-fifth the potency of phenobarbital (Dalton et al., 2003). Similarly, PXR activation by conazole fungicides is supported by activity in the PXR expression system and from toxicogenomic profile-like signatures of prototype PXR ligands (Goetz et al., 2006; Tully et al., 2006). Similar approaches have defined transcriptional mechanisms for pesticide induction of other CYPs through their activity as ligands for relevant transcriptional activators. The Ah receptor is known to be involved in the induction of Cyp isoforms in mice by MDP and related compounds (Cook and Hodgson, 1985, 1986) and may be assumed to be involved in any induction of Cyp1a1 or CYP1A1. The PPAR receptor has been shown to be involved in the induction, in mice, of Cyp4a10 and Cyp4a14 (Upham et al., 2007).

40.3  Inhibition As previously indicated, inhibition of XMEs can cause either an increase or a decrease in toxicity. Inhibitory effects can be demonstrated in a number of ways at different organizational levels.

40.3.1  Types of Inhibition and Experimental Demonstration 40.3.1.1  In Vivo Symptoms The measurement of the effect of an inhibitor (or inducer) on the duration of action of a drug in vivo is a common method of demonstrating its action, and previously the effects on the hexobarbital sleeping time or the zoxazolamine paralysis time were often used. Both of these drugs are fairly rapidly deactivated by the hepatic microsomal

Chapter | 40  Metabolic Interactions of Pesticides

949

cytochrome CYP-dependent monooxygenase system; thus, inhibitors of the CYP isoform(s) involved prolonging their action, whereas inducers have the opposite effect. Although, as a consequence of the availability of single, expressed isoforms for direct studies of inhibitory mechanisms, these methods are now used much less often, they are still valuable for demonstrating the effect in the intact organism, a necessity for risk assessment. In the case of activation reactions, such as the activation of the insecticide azinphosmethyl to its potent anticholinesterase oxon derivative, a decrease in toxicity is apparent when rats are pretreated with the CYP inhibitor SKF-525A.

reduced by the same procedures, because the carbamylated enzyme is unstable and, in addition, the residual carbamate is diluted. Microsomal monooxygenase inhibitors that form stable inhibitory complexes with CYPs, such as SKF-525A, piperonyl butoxide, and other methylenedioxyphenyl compounds, amphetamine and its derivatives and organophosphorus pesticides (OPs) containing the P  S moiety, can be readily investigated in this way because the microsomes isolated from pretreated animals have a reduced capacity to oxidize many xenobiotics.

40.3.1.2  Distribution and Blood Levels

In vitro measurement of the effect of one xenobiotic on the metabolism of another is by far the most common type of investigation of interactions involving inhibition. Although it is the most useful method for the study of inhibitory mechanisms, particularly when purified enzymes are used, it is of more limited utility in assessing the toxicological implications for the intact animal. The principal reason for this is that in vitro measurement does not assess the effects of factors that affect absorption, distribution, and prior metabolism, all of which occur before the inhibitory event under consideration and affect the concentration of inhibitor at the site of action. The primary considerations in studies of inhibition mechanisms are reversibility and selectivity. The inhibition kinetics of reversible inhibition give considerable insight into the reaction mechanisms of enzymes and, for that reason, have been well studied. In general, reversible inhibition involves no covalent binding, occurs rapidly, and can be reversed by dialysis or by dilution. Reversible inhibition is usually divided into competitive inhibition, uncompetitive inhibition, and noncompetitive inhibition. Because these types are not rigidly separated, many intermediate classes have been described. Although enzyme kinetics of microsomal enzymes are intrinsically difficult to elucidate due to the fact that the enzymes are membrane bound and both substrate and inhibitor are frequently lipophilic, methods for analysis of kinetic data are available that simplify the determination of the type of inhibition. For example, see the methods used in the study of the irreversible, mechanismbased inhibition of estradiol metabolism in humans by chlorpyrifos (Usmani et al., 2006). Competitive inhibition is usually caused by two substrates competing for the same active site. Following classical enzyme kinetics, there should be a change in the apparent Km but not in Vmax. In microsomal monooxygenase reactions, type I ligands, which often appear to bind as substrates but do not bind to the heme iron, might be expected to be competitive inhibitors, and this frequently appears to be the case. Uncompetitive inhibition has seldom been reported in studies of xenobiotic metabolism. It occurs when an inhibitor

Treatment of an animal with an inhibitor of xenobiotic metabolism may cause changes in the blood levels of an unmetabolized toxicant and/or its metabolites. This procedure may be used in the investigation of the inhibition of detoxication pathways; it has the advantage over in vitro methods of yielding results of direct physiological or toxicological interest because it is carried out in the intact animal. Moreover, the time sequence of the effects can be followed in individual animals, a factor of importance when inhibition is followed by induction. A refinement of this technique is to determine the effect of an inhibitor on the overall metabolism of a xenobiotic in vivo, by following the appearance of metabolites in the urine and feces and/or in blood or other tissue. Again, the use of the intact animal has practical advantages over in vitro methods, although little is revealed about the mechanisms involved.

40.3.1.3  Effects on In Vitro Metabolism Following In Vivo Treatment This method of demonstrating inhibition is of variable utility. The preparation of enzymes from animal tissues usually involves considerable dilution with the preparative medium during homogenization, centrifugation, and resuspension. As a result, inhibitors not tightly bound to the enzyme in question are lost, either in whole or in part, during the preparative processes. Therefore, negative results can have little utility because failure to inhibit and loss of the inhibitor give identical results. Positive results, on the other hand, not only indicate that the compound administered is an inhibitor but also provide a clear indication of excellent binding to the enzyme, most probably due to the formation of a covalent or slowly reversible inhibitory complex. The inhibition of acetylcholinesterase following treatment of the animal with organophosphorus compounds, such as paraoxon, is a good example, because the phosphorylated enzyme is stable and is still inhibited after the preparative procedures. In contrast, inhibition by carbamates is greatly

40.3.1.4  In Vitro Effects

950

interacts with an enzyme–substrate complex but cannot interact with free enzyme. Both Km and Vmax change by the same ratio, giving rise to a family of parallel lines in a Lineweaver– Burke plot. Noncompetitive inhibitors can bind to both the enzyme and enzyme–substrate complex to form either an enzyme– inhibitor complex or an enzyme–inhibitor–substrate complex. The net result is a decrease in Vmax but no change in Km. Metyrapone, a well-known inhibitor of monooxygenase reactions, can also, under some circumstances, stimulate metabolism in vitro. In either case, the effect is noncompetitive in that the Km does not change whereas Vmax does, decreasing in the case of inhibition and increasing in the case of stimulation. Irreversible inhibition, which is much more important toxicologically, can arise from various causes. In most cases, the formation of covalent or other stable bonds or the disruption of the enzyme structure is involved. In these cases, the effect cannot be readily reversed in vitro by either dialysis or dilution. The formation of stable inhibitory complexes may involve the prior formation of a reactive intermediate that then interacts with the enzyme (“suicide” or mechanism-based inhibition). An excellent example of this type of inhibition is the effect of the insecticide synergist piperonyl butoxide on hepatic microsomal monooxygenase activity, reviewed by Hodgson and Levi (1998) and Hodgson (1999). This methylenedioxyphenyl compound can form a stable inhibitory complex that blocks CO binding to CYP and also prevents substrate oxidation. This complex results from the formation of a reactive intermediate, and the type of inhibition changes from competitive to irreversible as metabolism, in the presence of NADPH and oxygen, proceeds. It appears probable that the metabolite in question is a carbene formed spontaneously by elimination of water following hydroxylation of the methylene carbon by the cytochrome (Dahl and Hodgson, 1979). Piperonyl butoxide inhibits the in vitro metabolism of many substrates of the monooxygenase system, including aldrin, ethylmorphine, aniline, aminopyrene, carbaryl, biphenyl, hexobarbital, and p-nitroanisole. The inhibition, by organophosphorus compounds such as ethyl p-nitrophenol thio-benzene phosphonate (EPN), of the carboxylesterase that hydrolyzes malathion is a further example of xenobiotic interaction resulting from irreversible inhibition, because in this case the enzyme is phosphory­ lated by the inhibitor. Oxons, such as chlorpyrifos oxon, are potent inhibitors of the esterases, in humans, that hydrolyze pyrethroids such as permethrin (Choi et al., 2004). Oxidative desulfuration of phosphorothioate pesticides such as chlorpyrifos, parathion, and fenitrothion by CYPs is known to release atomic sulfur, which covalently binds to and inactivates CYPs (Halpert et al., 1980; Kamataki and Neal, 1976; Levi et al., 1988; Neal and Halpert, 1982). In a recent study, administration of fenitrothion at a dose as low as 7 mg/kg inhibited the hydroxylation of 17 -estradiol by

Hayes’ Handbook of Pesticide Toxicology

hepatic microsomes (Berger and Sultatos, 1996). Usmani et al. (2003, 2006) have shown that chlorpyrifos and other OPs containing the P  S moiety are potent inhibitors of the human metabolism of both testosterone and estradiol. These results suggest that, in low concentrations, organophosphorus insecticides have the potential to inhibit enzymes important in normal sexual development. Another class of irreversible inhibitors of toxicological significance consists of those compounds that bring about the destruction of the xenobiotic-metabolizing enzymes. The drug allylisopropylacetamide, as well as other allyl compounds, has long been known to cause the breakdown of CYP and the resultant release of heme; the hepatocarcinogen vinyl chloride has also been shown to have a similar effect, probably also mediated through the generation of a highly reactive intermediate.

40.3.2  Synergism and Potentiation The terms synergism and potentiation have been variously used and defined but, in any case, involve a toxicity that is greater when two compounds are given simultaneously or sequentially within a short time frame than would be expected from a consideration of the toxicities of the compounds given alone. An example of synergism has already been mentioned. Piperonyl butoxide, sesamex, and related compounds increase the toxicity of insecticides to insects by inhibiting insect CYP. Other insecticide synergists that interact with CYP include aryloxyalkylamines such as SKF-525A, Lilly 18947, and their derivatives, compounds containing acetylenic bonds such as aryl-2-propynyl phosphate esters containing propynyl functions, phosphorothionates, benzothiadiazoles, and some imidazole derivatives. Insecticide synergists have similar interactions with mammalian CYPs. The best known example of potentiation involving insecticides and an enzyme other than the monooxygenase system is the increase in the toxicity of malathion to mammals that is brought about by certain other organophosphates. Malathion has a low mammalian toxicity due primarily to its rapid hydrolysis by a carboxylesterase. EPN, another organophosphorus insecticide, causes a dramatic increase in malathion toxicity to mammals at dose levels that, given alone, cause essentially no inhibition of cholinesterase. In vitro studies have shown that the oxygen analogue of EPN, as well as oxons of many other organophosphate compounds, increases the toxicity of malathion by inhibiting the carboxylesterase responsible for its degradation.

40.3.3  Antagonism In toxicology, antagonism may be defined as that situation in which the toxicity of two or more compounds administered together, or sequentially within a short time frame, is

Chapter | 40  Metabolic Interactions of Pesticides

less than would be expected from a consideration of their toxicities when administered individually. Apart from the effects mediated through induction of XMEs (discussed previously), antagonism does not appear to be important in pesticide interactions.

40.3.4  Pesticides as Inhibitors Examples of pesticides as inhibitors of the metabolism of other pesticides, other xenobiotics, or endogenous metabol­ ites in humans are shown in Table 40.3. Pesticides may act to inhibit CYPs or other enzymes by any of the mechanisms discussed previously – competitive inhibition, noncompetitive inhibition, or irreversible inhibition. As discussed more fully in the following section, MDP compounds have very complex interactions with the CYP system, being both inhibitors and inducers. Because MDP compounds are substrates for CYP enzymes, they may act initially as competitive inhibitors. As the MDP compound is metabolized, it becomes a suicide inhibitor with its reactive metabolite bound to the heme iron of CYP (Goldstein et al., 1973; Hodgson et al., 1998; Hodgson and Philpot, 1974). The herbicide synergist tridiphane, a postemergent herbicide, owes its activity to its ability to inhibit glutathione S-transferases. It has also been shown to induce epoxide hydrolase and CYP, specifically CYP4A, and peroxisomal enzymes (Levi et al., 1992; Moody and Hammock, 1987). In addition to induction of CYP4A, tridiphane functions as a selective CYP inhibitor, inhibiting CYP2B10 while having little or no effect on other CYP isoforms (Moreland et al., 1989). As assessed by in vitro studies, tridiphane appears to be a competitive inhibitor of CYP; its effect in vivo, however, is not yet known. Organophosphorus insecticides such as chlorpyrifos, parathion, and others that contain the P  S moiety are metabolized by the CYP system to the corresponding oxon, P  O, by oxidative desulfuration. This activation reaction, which converts the relatively inactive compound to a potent cholinesterase inhibitor, is thought to involve the formation of a P—S—O (phosphooxythirane) ring intermediate. Studies with both microsomes and purified enzymes (Halpert et al., 1980; Kamataki and Neal, 1976; Morelli and Nakatsugawa, 1978) have demonstrated that, during oxidative desulfuration, the released sulfur exists as a highly reactive molecule that then binds to the heme iron of CYP, inactivating the enzyme. This binding of reactive sulfur to CYP is accompanied by loss of CYP as detected by measurement of the dithionite-reduced CO complex as well as loss of monooxygenase activity (Berger and Sultatos, 1996; Butler and Murray, 1993; Neal, 1985; Neal and Halpert, 1982; Neal et al., 1983). Cohen (1984) showed that acetaminophen toxicity was reduced by the organophosphorus insecticide fenitrothion, as a result of inhibition of the CYP-dependent activation of acetaminophen. Studies

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with purified CYP isoforms and fenitrothion demonstrated that the amount of inhibition varied with the CYP isoform (Levi et al., 1988). In human liver microsomes, metabolism of parathion resulted in a concurrent loss of total CYP as well as the loss of several CYP-mediated enzyme activities (Butler and Murray, 1997). The activities inhibited included testosterone 6-hydroxylation, catalyzed by CYP3A4, 7-ethoxyresorufino-deethylation, catalyzed by CYP1A2, and tolbutamide methyl hydroxylation, catalyzed by CYP2C9/10. Aniline 4-hydroxylation, catalyzed by CYP2E1, was not inhibited. The inhibition of CYP-dependent monooxygenations by organophosphorus insecticides may be of considerable importance in human health risk assessment, as it has been shown that organophosphorus insecticides containing the P  S moiety are potent inhibitors of the human microsomal metabolism of both testosterone and estradiol (Usmani et al., 2003, 2006). Organophosphorus compounds may also inhibit enzymes other than CYP, particularly esterases (e.g., Cohen, 1984; Gaughan et al., 1980). Gaughan et al. (1980) also showed that profenofos, EPN, and S,S,S-tributylphosphorotrithioate (DEF), when administered in vivo to mice, all inhibited the liver microsomal esterases hydrolyzing trans-permethrin as well as the carboxylesterase hydrolyzing malathion. Chlorpyrifos oxon is a potent inhibitor of the hydrolysis of pyrethroids by human liver preparations (Choi et al., 2004). The fungicide captan, apparently through reactive metabolites, inhibits several CYP-dependent enzyme activities in mouse liver (Paolini et al., 1999) although it induces the 2-hydroxylation of testosterone. Methoxychlor, again through a reactive intermediate, inhibits the oxidation of both testosterone and estradiol, the pattern of metabolites indicating inhibition of CYP2C11 in rats and CYP3A in humans (Li et al., 1993). It may also be noted that nonpesticidal inhibitors of CYP isoforms may affect the metabolism and toxicity of pesticides. For example, Agyeman and Sultatos (1998) showed that the H2-blocker cimetidine caused a moderate increase in the toxicity of parathion but did not affect the toxicity of paraoxon, an effect brought about by the inhibition of CYP isoforms.

40.4  Biphasic effects: inhibition and induction Many inhibitors of mammalian monooxygenase activity can act also as inducers. Generally, inhibition of microsomal monooxygenase activity is fairly rapid and involves a direct interaction with the cytochrome, whereas induction is a slower process. Therefore, following a single injection an initial decrease due to inhibition is followed by an inductive phase. As the compound and its metabolites are eliminated, the levels of activity return to control values.

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Table 40.3  Inhibition of Human Hepatic Phase I Metabolism by Pesticides Substrate

Enzyme

Inhibitor(s)

Reference

Carbaryl

Liver microsomes

Chlorpyrifos

Tang et al. (2002)

Carbaryl

CYP2B6

Chlorpyrifos

Tang et al. (2002)

Carbofuran

Liver microsomes

Chlorpyrifos

Usmani et al. (2004)

DEET

Liver microsomes

Chlorpyrifos

Usmani et al. (2002)

Fipronil

Liver microsomes

Chlorpyrifos

Joo et al. (2007)

Fipronil

CYP3A4

Chlorpyrifos

Joo et al. (2007)

Imipramine

Liver microsomes

Chlorpyrifos, azinphosphos methyl, parathion

Di Consiglio (2005)

Imipramine

CYPs 1A2, 3A4, 2C19

Chlorpyrifos, azinphosphos methyl, parathion

Di Consiglio (2005)

Nonane

Liver microsomes

Chlorpyrifos

Joo et al. (2007)

Nonane

CYP2B6

Chlorpyrifos

Joo et al. (2007)

Estradiol

Liver microsomes

Chlorpyrifos, fonofos, carbaryl, naphthalene

Usmani et al. (2005)

Estradiol

CYP1A2

Chlorpyrifos, fonofos, carbaryl, naphthalene

Usmani et al. (2005)

Estradiol

CYP3A4

Chlorpyrifos, fonofos, deltamethrin, permethrin

Usmani et al. (2005)

Testosterone

Liver microsomes

Chlorpyrifos, phorate, fonofos

Usmani et al. (2003)

Testosterone

CYP3A4

Chlorpyrifos

Usmani et al. (2003)

Xenobiotic substrates

Endogenous substrates

Some of the best examples of compounds showing such biphasic effects are MDP compounds, such as the pesticide synergists piperonyl butoxide and sesamex, and the secondary plant compounds safrole and isosafrole. The effect of MDP compounds on CYP activity is an initial inhibition of activity followed by an increase above control levels (Kamienski and Murphy, 1971; Kinsler et al., 1990; Philpot and Hodgson, 1971–1972). The inhibitory effect of MDP compounds has been attributed to the formation of a stable inhibitory metabolite complex between the heme iron of the CYP and the carbene species formed when water is cleaved from the hydroxylated methylene carbon of the MDP compound (Dahl and Hodgson, 1979). Because CYP combined with MDP compounds in an inhibitory complex cannot interact with CO, the cytochrome CYP titer, as determined by the method of Omura and Sato (1964) (dependent upon CO binding to reduced cytochrome), reflects the biphasic effect. MDP exposure induces several hepatic CYP isoforms in mice, including CYP1A1, CYP1A2, and CYP2B10

(Adams et al., 1993a,b; Cook and Hodgson, 1985, 1986; Lewandowski et al., 1990; Ryu et al., 1995, 1996, 1997; Ryu and Hodgson, 1999). A number of studies have been published regarding the effects of MDP compounds on mammalian liver enzymes (for reviews see Adams et al., 1995; Hodgson et al., 1995a,b). It is apparent from extensive reviews of the induction of monooxygenase activity by xenobiotics that many compounds other than methylenedioxyphenyl compounds have the same biphasic effect. It may be that any synergist that functions by inhibiting microsomal monooxygenase activity could also induce this activity on longer exposure, resulting in a biphasic curve as described previously for methylene-dioxyphenyl compounds. This curve has been demonstrated for NIA 16824 (2-methylpropyl-2-propynyl phenylphosphonate) and WL 19255 (5,6-dichloro-1,2,3-benzothiadiazole), although the results were less marked with R05-8019 [2,(2,4,5-trichlorophenyl)-propynyl ether] and MGK 264 [N-(2-ethylhexyl)-5-norbornene-2,3-dicarboximide].

Chapter | 40  Metabolic Interactions of Pesticides

40.5  Activation Activation, as distinct from induction, is a stimulatory effect on enzyme activity caused by an interaction at the active site of the enzyme and/or an allosteric effect on enzyme protein conformation. As a consequence, activation tends to be rapid. Induction (Section 40.2), on the other hand, involves the synthesis of new enzyme and tends to be slower than activation. Although activation of CYP enzyme activity is less frequently encountered, and less well understood, than either inhibition or induction, it has been known for some time. Enhancement, by acetone, of the hepatic microsomal p-hydroxylation was first reported in 1968 (Anders, 1968). Flavone and benzoflavone both stimulate benzo(a)pyrene metabolism by rabbit liver CYPs, the extent of stimulation depending on the CYP isoform involved (Huang et al., 1981). 6-Hydroxylation of testosterone by the human isoform CYP3A4 is significantly increased by incubation of the enzyme with pyridostigmine bromide (Usmani et al., 2003), and Buratti and Testai (2007) have presented evidence for the autoactivation of CYP3A4 during dimethoate metabol­ ism. More recently, Cho et al. (2007) have shown that chlorpyrifos oxon significantly activates the production of 1-naphthol, 2-naphthol, trans-1,2-dihydronaphthalenediol, and 1,4-naphthoquinine from naphthalene by human liver microsomes. Further, it was shown that production of naphthalene metabolites by CYPs 2C8, 2C9, 2C19, 2D6, 3A4, 3A5, and 3A7 was activated by chlorpyrifos oxon while the production of naphthalene metabolites by CYPs 1A1, 1A2, 1B1, and 2B6 was inhibited by chlorpyrifos oxon. Activation effects on CYP metabolism of the insect repellent DEET (N,N-diethyl-m-toluamide) were also noted (Cho et al., 2007). Chlorpyrifos oxon inhibited the formation of N,N-diethyl-m-hydroxymethylbenzamide from DEET by human liver microsomes while stimulating the formation from DEET of N-ethyl-m-toluamide. This was reflected by the finding that CYP2B6, the principal isoform for N,N-diethyl-m-hydroxymethylbenzamide production, was inhibited by chlorpyrifos oxon while CYP3A4, the principal isoform for N-ethyl-m-toluamide production, was activated.

40.6  Hepatotoxicity Hepatotoxicity has frequently been observed as a consequence of xenobiotic exposure. Although XMEs and metabolic interactions may not be directly involved in hepatotoxicity, some consideration should be given to this phenomenon since loss of liver function may well give rise to consequences and interactions not seen in the intact liver. There have been a number of studies of enzyme induction in isolated hepatocytes from both surrogate animals and humans (see Section 40.2.1), but studies involving toxic

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effects of pesticides on the hepatocytes of surrogate animals have been relatively rare. The availability of the liverderived HepG2 cell line and the more recent availability of human hepatocytes have, however, made new approaches to this phenomenon possible. Typically, cell viability and cytotoxicity are measured by the use of the trypan blue exclusion method and the release of adenylate kinase into the medium. Apoptosis, or programmed cell death, is examined by measuring the induction of caspase3/7 (see Das et al., 2006 for a description of these methods). In both HepG2 cells and human hepatocytes the following pesticides caused cell death, release of adenylate kinase, and induction of caspase: fipronil and fipronil sulfone (Das et al., 2006); delta­ methrin and permethrin (Das et al., 2008a); DEET (Das et al., 2008b); and chlorpyrifos (Das et al., 2008b). The most potent of these is fipronil, while the least potent is DEET. Since fipronil sulfone is more active than fipronil, the CYP-dependent monooxygenation of fipronil may be considered an activation reaction.

Conclusion The conclusions reached in the preparation of the previous edition of this handbook remain viable, and considerable progress has been made in the realization of the anticipated outcomes. Knowledge of the metabolism of pesticides is essential and further knowledge is still needed for several reasons, including the development of more selective insecticides and for providing, in part, the fundamental basis for science-based risk assessments for human and environmental health. Since multiple exposures tend to be the rule rather than the exception, knowledge of metabolic interactions is a vital adjunct to the risk analysis process, one that is still inadequately understood or considered. Until relatively recently, and as a matter of necessity, this research was carried out almost exclusively on experimental animals, and the results, particularly in the case of human health risk assessments, were extrapolated to humans. Although much essential background will continue to be obtained from experimental animals, due to the new techniques of molecular biology and the availability of human cells, human cell fractions, and recombinant human enzymes, it is now possible to work directly on human biotransformations and metabolic interactions. Molecular techniques also permit the study of genetic polymorphisms that will enable us to identify populations at increased risk and enable studies to be carried out at the level of specific isoforms of the XMEs involved. The interaction of pesticides and clinical drugs, although long a subject for speculation, has been the subject of little investigation. The work of Di Consiglio et al. (2005) on the interaction between imipramine and organophosphorothionates makes it clear that much more work is needed on this problem.

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Thus the study of pesticide metabolism has entered a new molecular era that will be fascinating as well as useful.

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Chapter | 40  Metabolic Interactions of Pesticides

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Chapter | 40  Metabolic Interactions of Pesticides

Wattenberg, L. W. (1971). Studies of polycyclic hydrocarbon hydroxylases of the intestine possibly related to cancer. Effect of diet on benzpyrene hydroxylase activity. Cancer (Philadelphia) 28, 99–102. Welch, R. M., and Harrison, Y. (1966). Reduced drug toxicity following insecticide treatment. Pharmacologist 8, 217. Welch, R. M., Harrison, Y., and Burns, J. J. (1967). Hemorrhagic crises in the dog caused by dicumarol–phenobarbital interaction. Fed. Proc., Fed. Am. Soc. Exp. Biol. 26, 568. Welch, R. M., Levin, W., Kuntzman, R., Jacobson, M., and Conney, A. H. (1971). Effect of halogenated hydrocarbon insecticides on the metabolism and uterotropic action of estrogens in rats and mice. Toxicol. Appl. Pharmacol. 19, 234–246. Wilkinson, C. F., and Denison, M. S. (1982). Pesticide interactions with bio-transformation systems. In “Effects of Chronic Exposure to Pesticides on Animal Systems” (J. E. Chambers and J. D. Yarbrough, eds.), pp. 1–24. Raven Press, New York.

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Wright, A. S., Potter, D., Wooder, M. F., Donninger, C., and Greenland, R. D. (1972). Effects of dieldrin on mammalian hepatocytes. Food Cosmet. Toxicol. 10, 311–322. Wyde, M. E., Bartelucci, E., Ueda, A., Zhang, H., Yan, B. F., Negishi, M., and You, L. (2003). The environmental pollutant 1,1-dichloro-2,2-bis (p-chlorophenyl) ethylene induced rat hepatic cytochrome P4502B and 3A expression through the constitutive androstane receptor and pregnane X receptor. Mol. Pharmacol. 64, 474–481. Yu, T. L., and Varma, D. R. (1982). Pharmacokinetics, metabolism and disposition of salicylate in protein-deficient rats. Drug Metab. Dispos. 10, 147–152. Zhao, B., Baston, D. S., Hammock, B., and Denison, M. S. (2006). Interaction of diuron and related substituted phenylureas with the Ah receptor pathway. J. Biochem. Mil. Toxicol. 20, 103–113.

Chapter 41

Pesticide Excretion Ernest Hodgson North Carolina State University, Raleigh, North Carolina

41.1  Introduction Although, in the time that has elapsed since the publication of the second edition of this handbook (Krieger, 2001), studies of the excretory mechanisms for pesticide excretion in vivo have received little attention, two aspects have continued to advance: cellular elimination and the use of urinary metabolites as biomarkers of pesticide exposure. Except in simple life forms, elimination of toxicants, including pesticides and their metabolites, is part of a specialized system that, in addition to elimination, maintains the balance of water, minerals, and other substances necessary for terrestrial life. Pesticides, again typical of toxicants in general, are taken up by the body in most cases because of their lipophilicity. Before elimination is possible, they must first be metabolized into a form simulating that used by the body for the elimination of endogenous compounds. In general, they are metabolized by phase I and phase II xenobiotic-metabolizing enzymes (XMEs) to conjugation products that are more polar and hence more hydrophilic than the parent compound and then excreted by either the renal or the hepatic route. Although similar anion and cation transport systems are found in both kidney and liver, they differ in the type of excretory products eliminated. The renal system eliminates molecules of molecular mass smaller than 400–500 Daltons, whereas the liver handles larger molecules. The molecular mass threshold between renal and biliary excretion varies with species (Hirom et al., 1972), although in most species there are excretory products that are excreted by both systems. In addition to excretion via the bile, highly lipophilic chemicals that are recalcitrant to metabolism may be excreted as the parent chemical by a number of alternate routes, although in terms of the overall excretion of toxicant (including pesticides) these are generally of minor importance compared to urine and bile. General aspects of excretion of toxicants and their metabolites may be found in Levi et al. (1997), Matthews (1994), Pritchard and James (1982), and Wallace and Tarloff (2008), and a summary of the overall process is shown in Figure 41.1. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

Whole body Vectorial Transport

Transport through circulatory system in association with binding proteins Passive diffusion, carrier-mediated uptake, active uptake, filtration in elimination organs

Organs of elimination

Biotransformation

Cellular site of elimination

Passive diffusion, active transport out of body

Figure 41.1  Processes involved in the vectorial transport of xenobio­ tics from the whole body point of origin to the specific site of elimination. From G. A. LeBlanc, Elimination of Toxicants, in A Textbook of Modern Toxicology (E. Hodgson, ed.). John Wiley and Sons, 2004.

The excretion of pesticides and their metabolites has not been extensively investigated, perhaps because the rate of excretion seldom appears to be a rate-limiting step in the ultimate expression of toxicity. However, urinary metabol­ ites continue to be utilized as biomarkers of exposure.

41.2  Renal function 41.2.1  Overall Aspects The kidneys are primarily organs of excretion, and elimination by the kidney accounts for most by-products of normal body metabolism. They are also the primary organs for excretion of polar xenobiotics and polar metabolites of lipophilic xenobiotics. A useful description of kidney structure and function has recently been published by Tarloff and Wallace (2008). The functional unit of the kidney, the nephron, is shown in Figure 41.2. 961

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962

Glomerular Capsule

Proximal Tubules

To the Urinary Bladder

Artery

Glomerulus

Vein

Capillary Bed

Loop of Henle

Figure 41.2  The nephron of the mammalian kidney. From G. A. LeBlanc, Elimination of Toxicants, in A Textbook of Modern Toxicology (E. Hodgson, ed.). John Wiley and Sons, 2004.

41.2.2  Glomerular Filtration Passive filtration of the blood plasma in the glomerulus, under the influence of the blood pressure generated by the heart, is the initial step in urine formation. All molecules small enough to pass through the glomerular pores (70–100 Å) appear in the ultrafiltrate; any molecule larger than these pores or bound to molecules larger than these pores will not appear in the ultrafiltrate.

41.2.3  Tubular Reabsorption Tubular reabsorption is the second major step in urine formation. Most of the reabsorption of solutes necessary for normal body function such as amino acids, glucose, and salts takes place in the proximal part of the tubule. This reabsorption may be active, as in the case of glucose, amino acids, and peptides, whereas water, chloride, and other ions are passively reabsorbed. Reabsorption of water and ions also occurs in the distal tubule and in the collecting duct. Reabsorption of xenobiotics is usually passive and controlled by the same principles that regulate their passage across any membrane. That is, lipophilic compounds cross cell membranes more rapidly than polar ones; hence, lipophilic toxicants will tend to be passively reabsorbed more than polar ones and, overall, elimination of polar toxicants and their polar metabolites will be facilitated.

41.2.4  Tubular Secretion Tubular secretion is another important mechanism for excretion of solutes by the kidney. Secretion across the wall of

the tubule is generally active, with two systems, one for the secretion of organic acids, including conjugates, and the other for the secretion of organic bases. Passive secretion may occur as a result of a process known as diffusion trapping. Un-ionized weak acids and bases pass across the membrane into the lumen of the tubule and, depending on the pH of the urine, one or the other may become ionized and unable to diffuse back across the lumen wall. Diffusion trapping is, of course, extremely sensitive to variations in urine pH, a factor that may be utilized to speed elimination of toxicants. For example, alkalinization of the urine by ingestion of bicarbonate speeds up the elimination of salicylate. Tubular secretion, and hence excretion, of organic anions has been known to be of importance in the excretion of certain pesticides for some time (Pritchard and James, 1982). 2,4-Dichlorophenoxyacetic acid (2,4-D) and 2,4,5-trichlorophenoxyacetic acid (2,4,5-T) are usually applied as salts or esters, the latter being readily hydrolyzed in the body, and studies of their excretion have emphasized the parent acids, although various conjugates are also transported by the organic anion transport system (Erne, 1966; Pritchard and James, 1982). Active tubular secretion of 2,4-D has been demonstrated in a number of species, including rabbit (Dybing and Kolberg, 1967), rat (Fang et al., 1973), chicken (Erne and Sperber, 1974), dog (Hook et al., 1976), goat (Orberg, 1980), and flounder (Pritchard and James, 1979). 1 ,1,1-trichloro-2,2-bis( p-chlorophenyl)ethane (DDT) and its principal metabolite, 1,1-dichloro-2,2-bis (p-chlorophenyl)ethylene (DDE), are highly lipophilic and the latter is recalcitrant to further metabolism. Thus, DDT and, to a greater extent, DDE are sequestered in body lipids and have an extremely long half-life in the body. Some

Chapter | 41  Pesticide Excretion

Bile duct Hepatic Portal Vein

963

Canalicular Membrane

To Central Vein Hepatocytes

Sinusoid

Hepatic Artery

Figure 41.3  Diagrammatic representation of the basic architecture of the liver. From G. A. LeBlanc, Elimination of Toxicants, in A Textbook of Modern Toxicology (E. Hodgson, ed.). John Wiley and Sons, 2004.

portion of DDT, however, is metabolized to an organic acid, 2,2-bis(p-chlorophenyl) acetic acid (DDA), by dechlorin­ ation and oxidation at the 1-position (Pinto et al., 1965). DDA is a substrate for the organic acid transport system (Pritchard, 1976, 1978) and, as a consequence, is excreted considerably more rapidly than DDT or DDE.

41.3  Biliary Excretion Excretion by the liver, through the biliary system, has been known for a considerable time but, due to the difficulty in obtaining uncontaminated bile, has been less intensively investigated than renal excretion. A brief review of hepatic excretion may be found in Levi et al. (1997), and a representation of liver architecture is shown in Figure 41.3. Bile is secreted by the liver cells into the bile canaliculi. It then flows into the terminal branches of the bile duct, the hepatic duct, and the gallbladder. The contents of the gallbladder are discharged into the gut under the influence of hormones whose release is triggered by food ingestion. In species that lack a gallbladder, such as the rat, bile flows continuously into the duodenum. Secretion of xenobiotics or their metabolites into the bile is largely a function of molecular mass and may occur by passive diffusion or by active transport. Enterohepatic circulation is an important aspect of biliary excretion (Figure 41.4). Nonpolar xenobiotics are normally oxidized and then conjugated. If the molecular mass of the conjugate is appropriate for biliary excretion, it enters the gut where hydrolysis by intestinal microflora or gut conditions may occur. The compound, then being again in a less polar form, can be reabsorbed by the intestine and returned to the liver through portal circulation and the process repeated. Enterohepatic circulation thus increases the biologic half-life and possibly adverse effects of toxicants, particularly to the liver. For therapeutic purposes, the cycle can be interrupted by feeding an agent that binds the hydrolysis product and prevents its reabsorption, as in the use of cholestyramine in chlordecone poisoning.

Stomach

Liver

Gall bladder Bile Duct Intestines

Portal vein

Figure 41.4  Enterohepatic circulation.

41.4  Respiratory Excretion Volatile toxicants such as ethanol or pesticidal fumigants may be eliminated via the lungs, as may volatile metabolites, including acetone and carbon dioxide. Respiratory excretion is not known to be an important route for excretion of pesticides, in general, or their metabolites.

41.5  Other Routes of Excretion There are a number of other, less important routes of excretion, including sex-linked routes and alimentary elimination, and several routes based on natural secretory or growth processes are known.

41.5.1  Gender-Linked Routes of Excretion Certain routes of xenobiotic elimination are restricted to females, including excretion through milk, eggs, and fetus. Although such excretion is probably of minimal benefit to the mother, it may have serious consequences to the offspring.

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41.5.1.1  Milk Because milk is an emulsion of lipids in an aqueous protein solution, it may contain xenobiotics of many different physicochemical properties ranging from polar compounds such as alcohol and caffeine to less polar clinical drugs and to highly lipophilic chemicals such as DDT and DDE. Elimination of toxicants in milk is highly dependent on the biological half-life of the toxicant. Milk normally plays a minor role in the excretion of chemicals with short halflives but may be important for some chemicals with long half-lives. In experimental studies with chlorinated insecticides, up to 25% of the dose administered to cows was eliminated in the milk. In some South American countries, the DDT content of human mothers’ milk is close to the acceptable daily intake recommended by the World Health Organization (WHO). Although adverse effects on infants were not seen in these cases, when nursing mothers were accidentally exposed to hexachlorophene (Turkey) or polychlorinal biphenyls (PCBs) (Japan), signs of intoxication were seen in a number of infants.

41.5.1.2  Eggs Polar toxicants and metabolites may be eliminated in egg white and lipophilic compounds in the yolk. The effects of this on developing birds is controversial but may be significant, particularly if bioconcentration has occurred in the food chain. Effects of toxicants excreted into avian eggs should not be mistaken for the well-documented eggshell thinning, which is an effect on the female reproductive system.

41.5.1.3  Fetus The elimination of maternally derived toxicants in the fetus is of little or no benefit to the mother and, due to the generally small amounts involved, is usually of little or no harm to the fetus. However, as shown by the toxic effects of mercury, thalidomide, and diethylstilbestrol, this is not always the case.

41.5.2  Alimentary Elimination Passive elimination of lipophilic toxicants directly through the wall of the alimentary canal is probably, in most cases, unimportant, at least from a quantitative viewpoint. However, although slow, it may be an important route for excretion of chlordecone, particularly if reabsorption is prevented by administration of cholestyramine.

41.5.3  Obscure Routes of Excretion Because passive diffusion of lipophilic toxicants may occur across any cell membrane, it might be expected that

Hayes’ Handbook of Pesticide Toxicology

such chemicals will appear in many body secretions, such as sweat, or growth products, such as hair, nails, and skin. The sebaceous glands secrete an oily secretion and, probably for this reason, insecticides and PCBs have been found in human hair. Arsenic, mercury, and selenium have also been associated with hair. Although such routes of excretion are probably only a small proportion of the total excretion of any particular xenobiotic, they may provide a noninvasive method of estimating exposure or total body burden. Analysis of bird feathers is useful for the assessment of heavy-metal exposure, and the amount of cotinine, a major metabolite of nicotine, in saliva has been used extensively as a biomarker for nicotine uptake. The excretion of atrazine in saliva has also been tested in rats as a potential biomarker of exposure in exposed workers and shows promise for this ­application (Lu et al., 1997).

41.6  Cellular elimination To prevent concentration at toxic levels, hepatocytes and other cells have active transport processes to eliminate xenobiotics. Because the metabolism of xenobiotics generally yields products that are more polar and, consequently, have reduced capacity for passive diffusion compared to the parent compound, such transport processes are essential for cell viability. Since the publication of the second edition of the Handbook of Pesticide Toxicology (2001) there has been a dramatic increase in research on the role of transport proteins, or transporters, and, as a result, in our knowledge of their importance in the bioprocessing of xenobiotics (Miller, 2008). As yet this knowledge has not been applied extensively to pesticides; however, some general outlines on the role of transporters in pesticide bioprocessing are beginning to emerge. Many active transport proteins have been sequenced and classified into superfamilies and families. In mammals, those transporters with xenobiotic transport fall within the ABC superfamily, a group of ATP-dependent proteins (Miller, 2008). Of particular interest with regard to pesticides are p-glycoprotein and MRP (multi­drug resistance-associated protein). MRP has the capacity to transport glutathione, sulfate, or glucuronide conjugates, whereas p-glycoprotein is known to transport a wide array of xenobiotics, including some pesticides. Studies of the role of transporters in pesticide bioprocessing are generally of one of two types. In the first type the ability of the pesticide to inhibit the efflux of a chemical known to be transported by the transporter in question is measured, while in the second type the ability of the pesticide to bind to a particular transporter is measured. A useful summary of pesticides and transporters is included in Leslie et al. (2005). The importance of the role of transporters with regard to pesticides was first illustrated by the fact that p-glycoprotein

Chapter | 41  Pesticide Excretion

knockout mice died when treated with the miticide ivermectin, subsequently shown to be due to the accumulation of ivermectin in the brain due to the absence of p-­glycoprotein in the blood–brain barrier (Schinkel et al, 1994). This and subsequent studies (Lanning et al., 1996a, b; Macdonald and Gledhill, 2007) led to the conclusion that p-glycoprotein provided protection from the toxic effects of both ­ivermectin and chlorpyrifos. Other pesticides, including metolachlor (Leslie et al., 2001), fenitrothion, methoxychlor, and chlorpropham (Tribull et al., 2003), bind to human MRP1, and mrp1 knockout mice are more sensitive to the toxic effects of methoxychlor. Based on studies of a p-glycoprotein polymorphism, it has been suggested that p-glycoprotein plays a protective role in Parkinson’s disease (Drozdzik et al., 2003). On the basis of their ability to inhibit the p-­glycoproteinmediated efflux of doxorubicin, several pesticides of different chemical classes were shown to bind to human p-glycoprotein (Bain and LeBlanc, 1996). The most effective were the organochlorines, chlordecone, endosulfan, heptachlor, and heptachlor epoxide and the organophosphorus insecticides, chlorpyrifos, chlorthiophos, dicapthon, leptophos, parathion, and phenamiphos, as well as clotrimazole and ivermectin. None of the carbamates or pyrethroids tested was effective. Lipophilicity and molecular weight were major determinations of pesticide binding, with log Kow values of 3.6–4.5 and molecular weights of 391–490 Daltons being optimal. The authors point out that the ability to inhibit p-glycoprotein function does not necessarily mean that the chemical will be transported. Only endosulfan, the compound with the best binding characteristics, could be shown to be transported by p-glycogen. A study of four insecticides (Sreeramula et al., 2007), methylparathion, endosulfan, cypermethrin, and fenvalerate, demonstrated that all four stimulated p-glycoprotein ATPase activity at low concentrations. At higher concentrations the stimulation was lower or, in the case of methylparathion, inhibitory. It was further demonstrated that all of these insecticides inhibited the transport of a known ligand for p-glycoprotein, tetramethylrosamine. Other recent studies have shown that several pesticides were able to inhibit the uptake of a model p-glycoprotein substrate, calcein acetoxymethyl ester, into NIH 3T3 mouse fibroblasts stably transfected with the human MDR1 gene (Pivcevic and Zaja, 2006). Of the 14 pesticides tested, endosulfan, phosalone, and propioconzole were the most active. Similarly, rotenone, diazinon, and atrazine inhibited the efflux of taxol from the basolateral to the apical side of Caco-2 cells, and rotenone and diazinon inhibited estradiol-17-glucuronide uptake into MRP2expressed membrane vesicles while rotenone was a potent inhibitor of estradiol sulfate uptake into BCRP-expressed membrane vesicles (Pulsakar et al., 2006). The possibility that p-glycoprotein is related to insect resistance to insecticides (Buss et al., 2002; Buss and Callaghan, 2008; Lanning et al., 1996a, b) has not been fully explored and, although probable, remains hypothetical.

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41.7  Excretion of pesticides and their metabolites as biomarkers of exposure There have been a number of studies using urinary pesticides or their metabolites as biomarkers of exposures. The early studies in this area were summarized in 1989 (Wang et al, 1989). Some of these studies involve single compounds, primarily but not exclusively organophosphorus compounds, and examples are given in Table 41.1. Other studies are surveys of populations either exposed or potentially exposed to multiple pesticides. For example, urine from a sample of 1000 residents of the United States was analyzed for 12 analytes potentially derived from pesticides and 6 were frequently found (Hill et al., 1995). These, with possible parent compounds, were 2,5-dichlorophenol (from 1,4-dichlorobenzene); 2,4-dichlorophenol (from bifenox, clomethoxyfen, dichlofen-thion, etc.); 1-naphthol (from naphthalene, carbaryl, etc.); 2-naphthol (from naphthalene, etc.); 3,5,6-trichloro-2-pyridinol (from chlorpyrifos, chlorpyrifos-methyl); and pentachlorophenol (from pentachloro­ phenol, pentachloronitrobenzene). In another large study of multiple exposure, in tree nursery workers, only a small number, 42 out of 3134, of urine samples were positive, in this case for benomyl, bifenox, and carbaryl (Lavy et al., 1993). A summary of all methods, including measurement of urinary metabolites, of estimating exposure by biomarkers, was published in 2000 (Maroni et al., 2000), and appropriate analytical methods continue to be developed (e.g., for organophosphorus pesticides [OPs]) (De Alwis et al., 2008). A more recent study (McKone et al., 2007) used OP biomarker data to develop insights into the importance of different exposure sources in a cohort of almost 600. Gosselin et al. (2005) carried out a toxicokinetic modeling study of parathion and its metabolites (p-nitrophenol and alkyl phosphates) in humans in order to facilitate their use in exposure studies. Toxicokinetic modeling studies have also been carried out for chlorpyrifos and 2,4-D (Scher et al., 2008). Studies of exposure of schoolchildren and the excretion of pentachlorophenol (Wilson et al., 2007) and cis- and trans-permethrin (Morgan et al., 2007) showed good correlation in the case of permethrin, but an excess of excretion over estimated exposure in the case of pentachlorophenol indicated that the use of urinary biomarkers is not without problems and may need to be refined for future studies, depending upon the pesticide in question. Recently, urinary mercapturic acids have been extensively explored for use as biomarkers of exposure, and several detailed reviews are available (De Rooij et al., 1998; Van Welie et al., 1992). Although the emphasis has been on industrial and environmental chemicals, this potentially valuable technique has not been applied extensively to pesticides. However, the soil nematocide, dichloropropene, was included. Methods for the detection of pyrethroid insecticides, including pyrethrins, are being developed

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Table 41.1  Some Examples of the Use of Urinary Metabolites of Pesticides as Biomarkers of Exposure Pesticide

Urinary metabolite

Reference

Acephate

Methamidophos

Bouchard et al. (2006)

Carbaryl

1-Naphthol

Meeker et al. (2005, 2007)

Chlorpyrifos

Diethylphosphate, Diethylthiophosphate

Bouchard et al. (2005); Griffin et al. (1999)

3,5,6-Trichloro-2-pyrimidinol

Bouchard et al. (2005); Meeker et al. (2005)

Chlorpyrifos/quinalphos

Diethylphosphate, Diethylthiophosphate

Vasilic et al. (1992)

Chlorpyrifos-methyl

3,5,6-Trichloro-2-pyrimidinol alkylphosphates

Aprea et al. (1997)

Deltamethrin

3-Phenoxybenzoic acid, Cis-3-(2,2dimethylcyclo-propane-1-carboxylic acid

Ortiz-Perez et al. (2005)

Diazinon

1-Isopropyl-4-methyl-6-hydroxypyrimidine

Bouchard et al. (2006)

2,4-Dichlorophenoxyacetic acid (2,4-D)

2,4-D

Harris et al. (1992)

Dicofol

4,4’-Dichlorobenzilic acid

Nigg et al. (1991)

Guthion

Dimethylphosphorothioic acid

Franklin et al. (1981)

Malathion/thiometon

Dimethyl phosphate

Vasilic et al. (1999)

Dimethylphosphorothioate Dimethylphosphorodithioate Mono/dicarboxylic acids

Bouchard et al. (2006)

Naphthalene

1 and 2-Napthol

Meeker et al. (2007)

Organophosphorus insecticides

Alkyl phosphates

Azaroff (1999)

Permethrin (cis and trans)

3-Phenoxybenzoic acid

Morgan et al. (2007)

Pentachlorophenyl

Pentachlorophenyl

Wilson et al. (2007)

(Leng and Greis, 2005) and validated (Barr et al., 2007), and the use of biomarkers for pesticides other than OPs is being expanded. The effect of pesticides on the excretion of metabolites of endogenous metabolism has been explored to some extent. For example, in rats, treatment with dimethoate decreased the excretion of proline and lysine derivatives known to be collagen metabolites (Reddy et al., 1991). The mechanism of this effect was not investigated and the magnitude of the effect did not seem to be large enough for practical application. N-acetylglucosamidase was found to be slightly increased in the urine of applicators exposed to the soil nematocide, 1,3-dichloropropene, along with the principal metabolite of this nematocide, N-acetyl-S-(cis3-chloroprop-2-enyl) cysteine (Osterloh and Feldman, 1993).

Conclusion The excretion of pesticides in vertebrates, particularly mammals, is discussed. Renal and liver function, as they

relate to excretion, are summarized as well as the less important routes for excretion of pesticides: respiratory and alimentary. More obscure routes of excretion including gender-linked routes such as milk, eggs, placenta, and fetus as well as hair, sweat, etc., are briefly mentioned. The role of transporters in cellular elimination is considered, and, finally, the use of excreted pesticides and their metabolites as biomarkers of exposure is summarized.

References Aprea, C., Sciarra, G., Sartorelli, P., Sartorelli, E., Strambi, F., Farina, G. A., and Fattorini, A. (1997). Biological monitoring of exposure to chlorpyrifos-methyl by assay of urinary alkylphosphates and 3,5,6trichloro­pyridinol. J. Toxicol. Environ. Health 50, 581–594. Azaroff, L. S. (1999). Biomarkers of exposure to organophosphorus insecticides among farmers families in rural El Salvador: Factors associated with exposure. Environ. Res. 80, 138–147. Bain, L. J., and LeBlanc, G. A. (1996). Interaction of structurally diverse pesticides with the human MDR1 gene product p-glycoprotein. Toxicol. Appl. Pharmacol. 141, 288–298.

Chapter | 41  Pesticide Excretion

Barr, D. B., Leng, G., Berger-Preiss, E., Hoppe, H.-W., Weerasekera, G., Greis, W., Gerling, S., Perez, J., Smith, K., Needham, L. L., and Angerer, J. (2007). Cross validation of multiple methods for measuring pyrethroid and pyrethrum insecticide metabolites in human urine. Anal. Bioanal. Chem. 389, 811–818. Bouchard, M., Carrier, G., Brunet, R. C., Bonvalot, Y., and Gosselin, N. H. (2005). Determination of biological reference values for chlorpyrifos metabolites in human urine using a toxicokinetic approach. J. Occup. Environ. Hyg. 2, 155–168. Bouchard, M., Carrier, G., Brunet, R. C., Dumas, P., and Noisel, N. (2006). Biological monitoring of exposure to organophosphorus insecticides in a group of horticultural greenhouse workers. Ann. Occup. Hyg. 50, 505–515. Buss, D. S., McCaffery, A. R., and Callaghan, A. (2002). Evidence for p-glycoprotein modification of insecticide toxicity in mosquitoes of the Culex pipiens complex. Med. Vet. Entomol. 16, 218–222. Buss, D. S., and Callaghan, A. (2008). Interaction of pesticides with p-glycoprotein and other ABC proteins: a survey of the possible importance to insecticide, herbicide and fungicide resistance. Pestic. Biochem. Physiol. 90, 141–153. De Alwis, G. K. H., Needham, L. L., and Barr, D. B. (2008). Determination of dialkyl phosphate metabolites or organophosphorus pesticides in human urine by automated solid-phase extraction, derivatization and gas chromatography-mass spectrometry. J. Anal. Toxicol. 32, 721–727. De Rooij, B. M., Commandeur, J. N. M., and Vermeulen, N. P. E. (1998). Mercapturic acids as biomarkers of exposure to electrophilic chemicals: Applications to environmental and industrial chemicals. Biomarkers 3, 239–303. Drozdzik, M., Bialecka, M., Mysliwiec, K., Honczarenko, K., Stankiewicz, J., and Sych, Z. (2003). Polymorphism in the p-­glycoprotein drug transporter MDR1 gene: a possible link between environmental and genetic factors in Parkinson’s disease. Pharmacogenetics 13, 259–263. Dybing, F., and Kolberg, A. (1967). Inhibition of the renal tubular transport of p-aminohippurate (Tm-PAH) in the rabbit caused by subtoxic doses of dichlorophenoxyacetate (2,4-D). Acta Pharmacol. Toxicol. 25, 51–61. Erne, K. (1966). Distribution and elimination of phenoxyacetic acids in animals. Acta Vet. Scand. 7, 240–256. Erne, K., and Sperber, I. (1974). Renal tubular transfer of phenoxyacetic acids in the chicken. Acta Pharmacol. Toxicol. 35, 233–241. Fang, S. C., Fallin, E., Montgomery, M. L., and Freed, V. H. (1973). The metabolism and distribution of 2,4,5-trichlorophenoxyacetic acid in female rats. Toxicol. Appl. Pharmacol. 24, 555–563. Franklin, C. A., Fenske, R. A., Greenhalgh, R., Mathieu, L., Denley, H. V., Leffingwell, J. T., and Spear, R. C. (1981). Correlation of urinary pesticide metabolite excretion with estimated dermal contact in the course of occupational exposure to guthion. J. Toxicol. Environ. Health 7, 715–731. Gosselin, N. H., Bouchard, M., Brunet, R. C., Dumoulin, M. J., and Carrier, G. (2005). Toxicokinetic modelling of parathion and its metabolites in humans for the determination of biological reference values. Toxicol. Mech. Methods 15, 33–52. Griffin, P., Mason, H., Heywood, K., and Cocker, J. (1999). Oral and dermal absorption of chlorpyrifos: A human volunteer study. Occup. Environ. Med. 56, 10–13. Harris, S. A., Solomon, K. R., and Stephenson, G. R. (1992). Exposure of homeowners and bystanders to 2,4-dichlorophenoxyacetic acid (2,4-D). J. Environ. Sci. Health B 27, 23–38. Hill, R. H. Jr., Head, S. L., Baker, S., Gregg, M., Shealy, D. B., Bailey, S. L., Williams, C. C., Sampson, E. J., and Needham, L. L. (1995).

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Pesticide residues in urine of adults living in the United States: Reference range concentrations. Environ. Res. 71, 99–108. Hirom, P. C., Milburn, P., Smith, R. L., and Williams, R. T. (1972). Species variations in the threshold molecular-weight factor for the biliary excretion of organic acids. Biochem. J. 129, 1071–1077. Hook, J. B., Cardona, R., Osborn, J. L., Bailie, M. D., and Gehring, P. J. (1976). The renal handling of 2,4,5-trichlorophenoxyacetic acid (2,4,5-T) in the dog. Food Cosmet. Toxicol. 14, 19–23. Krieger, R. I., ed. (2001). “Handbook of Pesticide Toxicology,” 2 vols. Academic Press, San Diego, CA. Lanning, C. L., Fine, R. L., Corcoran, J. J., Ayad, H. M., Rose, R. L., and Abou-Donia, M. B. (1996a). Tobacco budworm p-glycoprotein: biochemical characterization and its involvement in pesticide resistance. Biochim. Biophys. Acta 1291, 155–162. Lanning, C. L., Fine, R. L., Sachs, C. W., Rao, U. S., Corcoran, J. J., and Abou-Donia, M. B. (1996b). Chlorpyrifos oxon interacts with the mammalian multidrug resistance protein p-glycoprotein. J. Toxicol. Environ. Health 47, 395–407. Lavy, T. L., Mattice, J. D., Massey, J. H., and Skulman, B. W. (1993). Measurements of year-long exposure to tree nursery workers using multiple pesticides. Arch. Environ. Contam. Toxicol. 24, 123–144. Leng, G., and Greis, W. (2005). Simultaneous determination of pyrethroid and pyrethrin metabolites in human urine by gas chromatography-high resolution mass spectrometry. J. Chromatog. B. Anal. Tech. Biomed. Life Sci. 814, 285–294. Leslie, E. M., Deeley, R. G., and Cole, S. P. (2001). Toxicological relevance of the multidrug resistance protein1, MRP1 (ABCC1) and related transporters. Toxicology 167, 3–23. Leslie, E. M., Deeley, R. G., and Cole, S. P. C. (2005). Multidrug resistance proteins: role of p-glycoprotein, MRP1, MRP2, and BCRP (ABCG2) in tissue defense. Toxicol. Appl. Pharmacol. 204, 216–237. Levi, P. E., Hodgson, E., and LeBlanc, G. A. (1997). Elimination of toxicants. In “A Textbook of Modern Toxicology” (E. Hodgson and P. E. Levi, eds.), 2nd ed. Appleton & Lange, East Norwalk, CT. Lu, C., Anderson, L. C., and Fenske, R. A. (1997). Determination of atrazine levels in whole saliva and plasma in rats: Potential of salivary monitoring for occupational exposure. J. Toxicol. Environ. Health 50, 101–111. Macdonald, N., and Gledhill, A. (2007). Potential impact of ABC1 ( p-glycoprotein) polymorphisms on avermectin toxicity in humans. Arch. Toxicol. 81, 553–563. Maroni, M., Colosio, C., Ferioli, A., and Fait, A. (2000). Introduction. Toxicology 143, 5–8; Organophosphorus pesticides. Toxicology 143, 9–37. Matthews, H. B. (1994). Excretion and elimination of toxicants and their metabolites In “Introduction to Biochemical Toxicology” (E. Hodgson and P. E. Levi, eds.), 2nd ed., Chap. 8. Appleton & Lange, East Norwalk, CT. McKone, T. E., Castorina, R., Harnly, M. E., Kuwabara, Y., Eskenazi, B., and Bradman, A. (2007). Merging models and biomonitoring data to characterize sources and pathways of human exposure to organophosphorus pesticides in the Salinas valley of California. Environ. Sci. Technol. 41, 3233–3240. Meeker, J. D., Barr, D. B., Ryan, L., Herrick, R. F., Bennett, D. H., Bravo, R., and Hauser, R. (2005). Temporal variability of urinary levels of nonpersistent pesticides in adult men. J. Exp. Anal. Environ. Epidemiol. 15, 271–281. Meeker, J. D., Barr, D. B., Serdar, B., Rappaport, S. M., and Hauser, R. (2007). Utility of 1-naphthol and 2-naphthol levels to assess environmental carbaryl and naphthalene exposure in an epidemiological study. J. Exp. Sci. Environ. Epidemiol. 17, 314–320.

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Miller, D. S. (2008). Cellular transport and elimination. In “Molecular and Biochemical Toxicology” (R. C. Smart and E. Hodgson, eds.). John Wiley and Sons, Hoboken, NJ. Morgan, M. K., Sheldon, L. S., Croghan, C. W., Jones, P. A., Chuang, J. C., and Wilson, N. K. (2007). An observational study of 127 preschool children at their homes and daycare centers in Ohio: environmental pathways to cis- and trans-permethrin exposure. Environ. Res. 104, 266–274. Nigg, H. N., Stamper, J. H., Deshmukh, S. N., and Queen, R. M. (1991). 4,4-Dichlorobenzilic acid urinary excretion by dicofol pesticide applicators. Chemosphere 22, 365–373. Orberg, A. (1980). Observations on 2,4-dichlorophenoxyacetic (2,4-D) excretion in the goat. Acta Pharmacol. Toxicol. 46, 78–80. Ortiz-Perez, M. D., Torres-Dorsal, A., Batres, L. E., Lopez-Guzman, O. D., Grimaldo, M., Carranza, C., Perez-Maldonado, I. N., PerezUrizar, J., and Diaz-Barriga, F. (2005). Environmental health assessment of deltamethrin in a malarious area Mexico: Environmental persistence, toxicokinetics, and genotoxicity in exposed children. Environ. Health Perspect. 113, 782–786. Osterloh, J. D., and Feldman, B. J. (1993). Urinary protein markers in pesticide applicators during a chlorinated hydrocarbon exposure. Environ. Res. 63, 171–181. Pinto, J. D., Camien, M. N., and Dunn, M. S. (1965). Metabolic fate of p,p-DDT [1,1,1-trichloro-2,2-bis(-chlorophenyl) ethane] in rats. J. Biol. Chem. 240, 2148–2157. Pivcevic, B., and Zaja, R. (2006). Pesticides and their binary combinations as p-glycoprotein inhibitors in NIH 3T3/MDR1 cells. Environ. Toxicol. Pharmacol. 22, 268–276. Pritchard, J. B. (1976). In vitro analysis of 2,2-bis(p-chlorophenyl) acetic acid (DDA) handling by rat kidney and liver. Toxicol. Appl. Pharmacol. 38, 621–630. Pritchard, J. B. (1978). Kinetic analysis of renal handling of 2,2-bis(pchlorophenyl) acetic acid by rat. J. Pharmacol. Exp. Ther. 205, 9–18. Pritchard, J. B., and James, M. O. (1979). Determinants of the renal handling of 2,4-dichlorophenoxyacetic by winter flounder. J. Pharmacol. Exp. Ther. 208, 208–286. Pritchard, J. B., and James, M. O. (1982). Metabolism and urinary excretion. In “Metabolic Basis of Detoxication: Metabolism of Functional Groups” (W. B. Jakoby, J. R. Bend, and J. Caldwell, eds.). Academic Press, San Diego. Pulsakar, S., Williams, D. A., LeDuc, B., Liu, N., and Xia, C. (2006). Interaction of structurally diverse pesticides with multidrug resistance proteins (P-GP, MRP2 and BCRP). Drug Metabol. Rev. 38 (suppl 2), 243.

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Reddy, P. N., Raj, G. D., and Dhar, S. C. (1991). Toxicological effects of an organophosphorus pesticide (dimethoate) on urinary collagen metabolites in normal and high protein diets fed female albino rats. Life Sci. 49, 1309–1318. Scher, D. P., Sawchuck, R. J., Alexander, B. H., and Adgate, J. L. (2008). Estimating absorbed dose of pesticides in a field study using biomonitoring data and pharmacokinetic models. J. Toxicol. Environ. Health A 71, 373–383. Schinkel, A. H., Smit, J. J. M., van Tellingen, O., Beijnen, J. H., Wagenaar, E., van Deemter, L., Mol, C. A. A. M., van der Valk, M. A., Runbanus-Maandag, E. C., te Riele, H. P. J., Berns, A. J. M., and Borst, P. (1994). Disruption of the mouse mdr1a p-glycoprotein gene leads to a deficiency in the blood–brain barrier and to increased sensitivity to drugs. Cell 77, 491–502. Sreeramula, A., Liu, R., and Sharom, F. J. (2007). Interaction of insecticides with mammalian p-glycoprotein and their effect on its transport function. Biochim. Biophys. Acta 1768, 1750–1757. Tarloff, J. B., and Wallace, A. D. (2008). Biochemical mechanisms of renal toxicity. In “Molecular and Biochemical Toxicology” (R. C. Smart and E. Hodgson, eds.), 4th ed. John Wiley and Sons, Hoboken NJ. Tribull, T. E., Bruner, R. H., and Bain, L. J. (2003). The multidrug resistance-associated protein 1 transports methoxychlor and protects the seminiferous epithelium from injury. Toxicol. Lett. 142, 61–70. Van Welie, R. T. H., van Dijck, R. G. J. M., and Vermeulen, N. P. E. (1992). Mercapturic acids, protein adducts, and DNA adducts as biomarkers of electrophilic chemicals. Crit. Rev. Toxicol. 22, 271–306. Vasilic, Z., Drevenkar, V., Rumenjak, V., Stengl, B., and Frobe, Z. (1992). Urinary excretion of diethylphosphorus metabolites in persons by quinalphos or chlorpyrifos. Arch. Environ. Contam. Toxicol. 22, 351–357. Vasilic, Z., Stengl, B., and Drevenkar, V. (1999). Dimethylphosphorus metabolites in serum and urine of persons poisoned by malathion or thiometon. Chem.-Biol. Interact. 119–120, 479–487. Wallace, A. D., and Tarloff, L. B. (2008). Biochemical mechanisms of renal toxicity. In “Molecular and Biochemical Toxicology” (R. C. Smart and E. Hodgson, eds.), 4th ed. Wiley, New York. Wang, R. G. M., Franklin, C. A., Honeycutt, R. C., and Reinert, J. C. (ed.) (1989). “Biological Monitoring for Pesticide Exposure: Mea­ surement, Estimation and Risk Reduction”. Americal Chemical Society, Washington, DC. Wilson, N. K., Chuang, J. C., Morgan, M. K., Lordo, R. A., and Sheldon, L. S. (2007). An observational study of the potential exposures of preschool children to pentachlorophenol, bisphenol A and nonylphenol at home and daycare. Environ. Res. 103, 9–20.

Section VII

Exposure Measurement and Mitigation

(c) 2011 Elsevier Inc. All Rights Reserved.

Chapter 42

Exposure Framework Linda S. Sheldon U.S. Environmental Protection Agency, National Exposure Research Laboratory, Research Triangle Park, North Carolina

42.1  Introduction Pesticides are substances used to repel, control, or kill certain forms of plant or animal life that are considered to be pests. Pesticides include herbicides for destroying weeds and other unwanted vegetation, insecticides for controlling a wide variety of insects, fungicides used to prevent the growth of molds and mildew, disinfectants for preventing the spread of bacteria, and compounds used to control mice and rats. Since nearly 80% of pesticides in the United States are agricultural chemicals used in food production, people are exposed to low levels of pesticide residues through their diets. Pesticides are also widely used in a variety of other settings including homes, schools, hospitals, and workplaces, which can lead to additional exposures. Scientists do not yet have a clear understanding of the health effects of exposures to these pesticide residues, although evidence suggests that children may be particularly susceptible to adverse effects. Because pesticides are designed to be toxic and because there is widespread low-level exposure, it is important to both understand and control exposures and risks while still maintaining the ­benefits of these chemicals. The Food Quality Protection Act (FQPA) provides the regulatory mandate for assessing and managing pesticide risks with a special emphasis on protecting children. FQPA requires the U.S. Environmental Protection Agency (EPA) to use exposure assessments in the pesticide tolerance setting process. These exposure assessments must consider the aggregate exposure and cumulative risks of infants and children to pesticides from all sources and by all routes. Implicitly, FQPA requires that risk assessments must be based on exposure data that are high quality and high quantity or exposure models using factors that are based on existing reliable data. Exposure science evaluates and predicts exposures and provides information for developing risk assessments Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

as well as the most effective strategies for reducing these risks. It describes the world we live in, the ways in which we interact with that world, and the potential consequences of our interactions. This is done by identifying and characterizing pesticides and human behaviors and the processes that drive exposures. The following section provides an overview of exposure science. It also provides information on unique considerations for pesticides, including the need to explicitly evaluate exposures for children in order to ensure their protection.

42.2  Exposure science Exposure is the contact of an individual with a contaminant for a specific duration of time (IPCS, 2004; Zartarian et al., 2005). For exposure to occur the contaminant and the individual must come together in both space and time. Exposure science characterizes and predicts this intersection (U.S. EPA, 1992). Barr (2006) has defined human exposure science as the study of human contact with chemical, physical, or biological agents occurring in their environments intended to advance knowledge of the mecha­nisms and dynamics of events either causing or preventing adverse health outcome. We have broadened this definition in order to include the important role of exposure science in effectively preventing or mitigating risks. Considering this expanded view, exposure science is defined as characterizing and linking the processes that impact the transport and transformation of contaminants from their source through human contact to target tissue dose. Exposure is described in terms of the magnitude, frequency, and duration of contact. For most contaminants, the magnitude of exposure is a critical characteristic in determining adverse effects. Likewise, both the frequency and the timing of exposures can have an important impact. Exposure can be either continuous or intermittent depending 971

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upon the source of the contaminant, its persistence in the environment, and activities of the individual that lead to contact with the contaminant. Additionally, for some contaminants, there are very specific life stages (such as fetal development) during which specific and characteristic exposure routes may predominate and where exposure will lead to an enhanced adverse outcome. Figure 42.1 is an adaptation of the source-to-outcome framework developed by the National Research Council (NRC, 1983, 1998). The processes that are important for exposure science start with a contaminant entering the environment and end with dose characterization. Starting in the upper left-hand corner, contaminants (primarily chemical or biological) are released into the environment from a source. Many contaminants can be transformed through a number of processes, including chemical reactions and biological degradation. Contaminants or their transformation products move through the environment and can be found in environmental media including air, water, soil, dust, and food. The intensity of exposure depends upon the concentration in the media, as well as the duration of ­contact with the receptor. Exposure becomes “dose” when the stressor moves across the receptor’s body barrier. The text under each box in Figure 42.1 shows the information that is used to characterize the various processes represented in the boxes. The arrows between the boxes represent models that are used to link the processes. Exposure to environmental contaminants is considerably more complex than illustrated in Figure 42.1. Multiple contaminants enter the environment at the same time from many different sources. Contaminants can remain

unchanged or they can be transformed by physical, chemical, or biological processes to become different agents. These contaminants or their transformation products can partition and move through many different environmental media (i.e., air, water, soil, sediment, and the plant and animal life of a particular region). Contaminants or their transformation products can come from several sources through a number of different pathways and routes. Sources include all uses of a chemical that could result in exposure. Route of exposure (i.e., dermal, oral, inhalation) is defined as the portal of entry. There are three routes of exposure: the skin is the portal of entry for the dermal exposure route, the mouth is the portal of entry for the ingestion exposure route, and the lung is the portal of entry for the inhalation exposure route. Pathway is defined as the course that the contaminant takes from its source to the portal of entry. Exposure factors are the factors related to human behavior and characteristics that determine an individual’s exposure to a contaminant. For example, an individual’s exposure to a pesticide by the inhalation route is determined by factors that include the duration of time spent in different microenvironments during the day and the individual’s inhalation rates during the period of exposure. Aggregate exposure is the sum of exposures to a single stressor from all sources and pathway(s) over a given time period. Cumulative risks are those that result from aggregate exposures to a single stressor over multiple time periods, or from concurrent and/or synergistic exposures to multiple stressors. Definitions that are pertinent to understanding exposure concepts are given in Table 42.1 (U.S. EPA, 2001).

Figure 42.1  Source-to-outcome framework for human health exposure research.

Chapter | 42  Exposure Framework

Table 42.1  Definitions Related to Pesticide Exposure Term

Definition

Acute exposure

An exposure period of less than 1 day

Aggregate exposure

The combined exposures to a single chemical from all sources across all routes and pathways

Chronic exposure

An exposure presumed to occur over a substantial portion of the individual’s lifetime

Cumulative exposure

The total exposure to chemicals that cause a common toxic effect(s) to human health by the same, or similar, sequence of major biochemical events

Exposure

The contact (at visible external boundaries) of an individual with a pollutant for specific durations of time

Exposure factors

The factors related to human behavior and characteristics that determine an individual’s exposure to a pesticide or contaminant. For example, duration of exposure, inhalation rates, transfer coefficients

Exposure pathway

The course that the chemical takes from its source to the receptor’s portal of entry

Exposure route

The portal of entry of a chemical into the body

Exposure scenario

The combination of facts, assumptions, and inferences that define a discrete situation or activity where potential exposures may occur. These include the source, the exposed population, the time frame of exposure, microenvironment(s), and activities

Intermediate-term exposure An exposure lasting from 1 week to several months Pathway

The course that the contaminant takes from its source to the portal of entry

Short-term exposure

An exposure lasting from 1 to 7 days

42.3  Exposure assessment Exposure assessment is the process for identifying potentially exposed populations and quantifying exposures. Essentially, exposure assessments seek to characterize

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“real-life” situations whereby (1) potentially exposed popu­ lations are identified, (2) potential pathways of exposure are identified, and (3) the magnitude, frequency, and duration of chemical intakes/potential doses are quantified. The adverse impact of exposure depends upon the characteristics of the exposure, the potency of the contaminant, and the susceptibility of the individual. The greatest adverse impact of any given contaminant will be to those individuals or populations that are most exposed and/or most susceptible to the exposure. This concept is illustrated in Figure 42.2. Vulnerability refers to characteristics of an individual or population that places them at increased risk of an adverse effect (U.S. EPA, 2003). The text box shows some of the ways that a receptor may be more vulnerable. Included are factors that can lead to increased susceptibility or higher exposure. Susceptibility refers to characteristics that lead to a greater response for the same exposure. The concepts of differential exposure and susceptibility are crucial if the goal is to protect not only the general population, but also those populations at greatest risk. Exposure assessments, therefore, should identify and understand those conditions that lead to the highest contaminant intensities and resulting exposures, as well as those situations that lead to exposure for the most susceptible populations. The desire to protect vulnerable populations is the concept behind the focus on children when considering pesticide exposure and risk. Exposure assessments may be conducted using either direct or indirect approaches. Direct assessments measure the contact of the person with the chemical concentration in the exposure media over an identified period of time. There are very few cases where methods exist and are used to make direct exposure assessments. Personal monitoring techniques such as the collection of personal air or duplicate diet samples are used to directly measure exposure to an individual during a point in time. Indirect assessments use available information on concentrations of chemicals in exposure media, along with information about when, where, and how individuals might contact the exposure media. The indirect approach then uses models and a series of exposure factors (i.e., pollutant concentration, contact duration, contact frequency) to estimate exposure. For a few pesticides, biomarkers can serve as a useful measure of direct exposure aggregated over all sources and pathways. It should be understood that biomarkers will measure integrated exposure from all routes. However, to use biomarkers for this purpose, several important criteria must be met. Biomarkers that can accurately quantitate the concentration of a pesticide or its metabolite(s) in easily accessible biological media (blood, urine, breath) must be identified and available. The pharmacokinetics of absorption, metabolism, and excretion must be known. Finally, the time between pesticide exposure and biomarker sample collection must be known. Although there are a number of biomarkers that meet these criteria, very few studies using

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Figure 42.2  The highest concentrations and the most susceptible populations create the greatest potential risk.

biomarkers have collected all of the information required to accurately estimate exposure. Exposure assessments can use models that are either deterministic or probabilistic. A deterministic model provides a point estimate of exposure. A probabilistic model considers the range of estimates and provides a probability distribution of exposures. As an example, a deterministic dietary exposure would assume that an individual eats a mass of food per day with a given concentration of a pesticide residue. A probabilistic exposure model would calculate the range of food mass eaten and the range of food types (with each food type having a given concentration of a pesticide residue) eaten by a particular subset of individual (e.g., of a certain age and gender). The result would be identification of the age and gender of individuals likely to be in the upper percentile of the exposure distribution. Probabilistic models can also be used to identify the number of individuals in a population likely to be at risk.

42.4  Considerations for pesticide exposure assessments As stated earlier, FQPA requires the EPA to develop exposure assessments when setting a food tolerance. There are several important aspects that are required as part of these assessments. First, these assessments must evaluate aggregate exposures for a single pesticide from all potential sources, including residues in the food in question (i.e., the commodity for which a tolerance is being sought), residues in other foods for which there are tolerances, residues in groundwater or surface water that is consumed as drinking water, and residues from other nondietary, nonoccupational uses of the pesticide (e.g., residential and other indoor/outdoor uses). Figure 42.3 demonstrates the complexity associated

with conducting aggregate exposure assessments. This figure graphically depicts all of the potential sources, routes, and pathways for children’s exposure to a pesticide. Such a conceptual model should serve as the starting point when developing aggregate exposure assessments for pesticides. Not only will a conceptual model provide the basis for identifying all sources and pathways, it will also provide the context for quantifying and prioritizing the most important sources and pathways. The EPA has developed standard operating procedures that provide specific procedure for characterizing the important exposure ­ scenarios for nonoccupational pesticide exposures (U.S. EPA, 1997, 2006). Cumulative risks due to aggregate exposures to all pesticides with a common mechanism of action must also be assessed. The science for accurately conducting cumulative risk assessments is still developing. The development and application of linked exposure and dose models are considered crucial for both aggregate and cumulative risk assessments for pesticides. These linked models provide the unique ability to examine exposure and target tissue dose characteristics for multiple pathways, multiple routes of exposure, and multiple pesticides simultaneously. To complete the risk assessment, outputs of exposure/dose models must also be linked to models of chemical toxicity. Exposure assessments require that not only the magnitude, but also the frequency and duration must be described. For pesticides, four exposure durations generally are considered. Acute exposure is defined as an exposure period of less than 1 day. Exposures through food and drinking water have been included in acute exposure assessments. Short-term exposure is defined as an exposure lasting from 1 to 7 days. Possible short-term exposures to pesticides in and around the home could come from uses such as on lawns, ornaments, and home gardens; as a crack and crevice treatment for insects; as a treatment for

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Respiratory Tract

Inhalation

Exposure Media

Source indoor residential outdoor residential industrial agricultural

Release & Transfer

Outdoor air water soil plants surfaces

Inhalation Rate Activity Patterns

Dermal Contact Mass Transfer Rates

Inhalation Exposure Rate

R.T. Uptake Uptake Rate

Skin Surface Dermal Exposure Rate

Dermal Uptake Uptake Rate

Absorbed Dose

Contact Activities Indoor air water house dust food surfaces clothes

Body

Mouthing Activities Nondietary Ingestion Ingestion Rate Mouthing Activities

Dietary Ingestion Ingestion Rate Diet

Ingestion Rate G.I. Uptake G.I. Tract

Uptake Rate

Ingestion Exposure Rate

Exposure Figure 42.3  Conceptual model of children’s residential exposure to pesticides.

carpets or other surfaces; and as a flea treatment for pets. Other short-term exposures could occur in public places such as parks, school playgrounds, and playing fields. Intermediate-term exposure is defined as exposure lasting from 1 week to several months. Possible ­intermediate-term exposures to pesticides in and around the home could occur due to use of rodenticides as well as some of the exposure scenarios described above in the acute and short-term categories. Chronic exposure is presumed to occur over a substantial portion of the individual’s lifetime. Although chronic exposure can occur via all routes and pathways, dietary is considered to be the largest ­ component. Pesticides that are used as termite control could also result in chronic exposures. An additional and unique requirement of FQPA is the need to consider special vulnerabilities of children, and specifically to conduct exposure and risk assessments for infants and young children. In evaluating environmental health risks to children, it is important to understand that children are not little adults. Children’s exposures to environmental contaminants and consumer products are expected to be different and, in many cases, much higher than older individuals (U.S. EPA, 2007). These differences

in exposure are due to differences in physiological function and surface-to-volume ratio. Children’s behavior and the way that they interact with their environment may have a profound effect on the magnitude of their chemical exposures. Children crawl, roll, and climb over contaminated surfaces, resulting in higher dermal contact than would be experienced by adults in the same environment. Children eat different foods that may result in higher dietary ingestion. Children’s mouthing activities (hand-to-mouth and objectto-mouth) will result in indirect ingestion of chemicals if the hands or objects are contaminated. Increased indirect ingestion of contaminants also occurs when ­children handle and eat foods that have come in contact with the floor or other contaminated surfaces. Again, Figure 42.3 is a conceptual model depicting sources, pathways, and routes of exposure for children (Cohen Hubal et al., 2000). It is important to understand that the activities listed above not only will result in differences in exposures between children and adults, but also will result in differences in exposures among children of different developmental stages. Thus, exposure assessments should be required for children in each age group, with age group being defined by the developmental stage of the child.

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Guidelines for conducting risk assessments for children have been developed along with the methods for quantifying specific exposure scenarios for nonoccupational pesticide uses.

Conclusion Pesticides provide many benefits to society, including improved food production and control of insects as disease vectors. However, these chemicals are designed to be toxic and, as a consequence, there may be risks associated with their use. In order to ensure the safe use of pesticides we must clearly understand the risks associated with these chemicals. Exposure science is fundamental to this process in that it provides the knowledge and tools to identify and quantify exposures from pesticides, to prioritize the importance of those exposures, and to design and evaluate options for reducing those exposures that lead to the highest risk. Understanding aggregate exposures to and cumulative risks of pesticides requires estimating simultaneous exposures and doses to multiple stressors from multiple routes and sources. Predictive exposure models can be used to understand exposures to either new pesticides or new applications of pesticides. The concepts and application of exposure science are still developing, especially when we consider exposures to young children. Currently, there are limited “real-world” exposure data. To collect these data, new, more efficient methods will be needed. In the future, models should provide the basic framework for conducting exposure research and using it in the context of regulatory decision making. However, we have a restricted mechanistic understanding of important processes that lead to exposure. Singleprocess models need to be improved and linked along the source-to-dose continuum. New biomonitoring methods and approaches for using and interpreting these data need to be developed. Finally, studies using exposure measurements and biomonitoring need to be conducted to improve and evaluate our understanding of the critical process associated with pesticide exposures.

References Barr, D. (2006). Human exposure science: a field of growing importance. J. Expo. Sci. Environ. Epidemiol. 16, 473. Cohen Hubal, E. A., Sheldon, L. S., Zufall, M. J., Furke, J. M., and Thomas, K. (2000). The Challenge of assessing children’s residential exposure to pesticides. J. Expo. Sci. Environ. Epidemiol. 10, 638. IPCS (International Programme on Chemical Safety) (2004). “Risk Assessment Terminology. Part 2: Glossary of Key Expsoure Assessment Terminology (Harmonization Project Document No. 1)”, World Health Organization, Geneva, Switzerland. NRC (National Research Council) (1983). “Risk Assessment in the Federal Government: Managing the Process,” The National Academies Press, Washington, DC. NRC (1998). “Research Priorities for Airborne Particulate Matter: I. Immediate Priorities and a Long-Range Research Portfolio,” The National Academies Press, Washington, DC. U.S. EPA. 2007. Important Exposure Factors for Children. An Analysis of Laboratory and Observational Field Data Characterizing Cumulative Exposure to Pesticides. Washington, DC: Office of Research and Development. EPA/600/R-07/013. U.S. EPA. 2001. Draft Protocol for Measuring Children’s NonOccupational Exposure to Pesticides by all Relevant Pathways. Research Triangle Park, NC: Office of Research and Development. EPA/600/R-03/026. U.S. EPA. 2006. A Framework for Assessing Health Risks of Environmental Exposure to Children. Washington, DC: US EPA. EPA/600/R-05/093F. http://cfpub.epa.gov/ncea/cfm/recordisplay. cfm?deid158363 U.S. EPA. 2003. Framework for Cumulative Risk Assessment. U.S. Environmental Protection Agency, Washington, DC, EPA/600/P02/001F, 2003. U.S. EPA. 1997. Standard Operating Procedures (SOPs) for Residential Exposure Assessments. Office of Prevention, Pesticides, and Toxic Substances, U.S. Environmental Protection Agency, Washington, DC. http://www.epa.gov/scipoly/sap/meetings/1997/september/sopindex.htm U.S. EPA. 1992. Guidelines for Exposure Assessment. Risk Assessment Forum, U.S. Environmental Protection Agency, Washington, DC. EPA/600/Z-92/001, May 1992. Zartarian, V., Bahadori, T., and McKone, T. (2005). Adoption of an Official ISEA Glossary. J Expo Anal Environ Epidemiol 15, 1.

Chapter 43

Sampling and Analysis for Nonoccupational Pesticide Exposure Assessments Roy Fortmann, Nicolle S. Tulve and M. Scott Clifton U.S. Environmental Protection Agency, National Exposure Research Laboratory, Research Triangle Park, North Carolina

43.1  Introduction Pesticides are used extensively in the United States to control a variety of pests. Commercial agriculture and nonagricultural industries account for about 80% of the total pesticide use in the United States, while the remaining 20% is used for pest control associated with home, garden, yard, and pets. Pesticides frequently occur in the indoor environments that we occupy as a result of indoor applications, spray drift and infiltration, and track-in from outdoor applications. To fully assess human exposures to pesticides, it is necessary to understand the aggregate exposures in all of the microenvironments that people occupy. As discussed in other chapters of this handbook, it is critical to accurately assess and predict exposures in order to perform risk assessments for pesticides. The Food Quality Protection Act of 1996 requires that both aggregate exposure and cumulative risk be assessed. Aggregate exposure considers the exposure to a pesticide by all routes and pathways (inhalation, dietary and nondietary ingestion, and dermal absorption). Cumulative risk considers the risk from aggregate exposures to all pesticides having a common mechanism of toxicity. The measurement of pesticides or pesticide metabolites in biological specimens (e.g., urine), known as biomonitoring, is an important tool used to evaluate human exposure. Modeling is another important tool used to support exposure assessments. An extensive array of models has been developed and is described in other chapters in the handbook. Despite the significant advances in biomonitoring and modeling tools, there is a continuing need to perform measurements of pesticides and pesticide metabolites in environmental media and to collect ancillary information to assess Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

exposures. These measurement studies are critical for determining the factors that have the most significant impact on people’s exposures and the spatial and temporal variability of exposures as people go about their normal activities in their everyday environments. Measurement studies that incorporate the normal activities of people in their everyday environments fall under the category of studies described as observational exposure studies. It is important to note that when conducting observational studies, there should be no attempt to control the activities or actions of the participants that could impact their exposures. Any attempt to do so introduces significant ethical concerns and introduces bias that prevents obtaining the very data the studies are designed to collect, which is exposure information based on normal activities in everyday environments. Various approaches are available for assessing human exposure to pesticides. Direct methods involve measurements of pesticides and/or their metabolites in environmental media, diet samples, and/or biological media. These data and selected ancillary information are used in simple algorithms, statistical methods, and models to estimate exposures. Indirect methods estimate individual’s exposures through the use of questionnaires, diaries, exposure surrogates, or deterministic algorithms. Probabilistic methods are used to develop population-level exposure estimates. This chapter focuses on the direct measurement methods used in observational human exposure studies and includes approaches employed in the design of observational studies and considerations that should be addressed during study design and implementation. State-of-the-science sample collection methods and considerations for the selection of methods used for estimating exposures associated with different 977

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routes and pathways of exposure are discussed. This discussion also includes the general principles and state-ofthe-science analytical methods used to measure pesticides and metabolites in environmental media, diet samples, and biological samples. Measurements are centered around estimating exposures of people in their real-world environments as they go about their normal day-to-day activities and apply to current-use pesticides applied indoors (e.g., crack and crevice treatments, sprays, foggers, baits, gels) by occupants or professional applicators in buildings such as residences, child care centers, schools, and public access buildings. Occupational exposures are not addressed, although some of the methods may be applicable if differences in exposure routes, concentrations, and timing of exposures are considered.

43.2  Design of nonoccupational observational exposure measurEment studies This section describes guidance for designing observational measurement studies to assess exposure concentrations and exposure factors in nonoccupational settings. Nonoccupational settings refer to locations such as single and multi-residential housing units and also include child care centers, schools, public access buildings, and other nonoccupational locations. The term “observational” is used to distinguish between measurements taken in everyday environments versus experiments conducted in a controlled laboratory system. Observational exposure measurement studies are performed for many different reasons. Examples include determining occurrence/co-occurrence and concentrations of pesticides in environmental media in a microenvironment (e.g., residence, school, public access building, yard, vehicle); identifying the important routes and pathways of exposure for different chemicals and chemical classes; determining which are the most important factors and activities affecting exposure; estimating exposure and dose for the exposed individual; evaluating exposure or dose models; and evaluating intervention and risk mitigation approaches and methods. A study may be designed to meet one or more study objectives, or it may be hypothesis driven (e.g., that diet is the primary route of exposure for the targeted pesticide). Regardless, the study objectives or hypothesis must be clearly defined in order to identify the data analyses and the data required to address the objectives. It is critical that the data analysis plan be developed during the study design phase because the data analysis plan should serve as the basis for the design of the study; it should not be developed in response to the data collected. Unlike worker exposure, which is discussed in other chapters of the handbook and for which there are several reviews and guidance documents on methodology for

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exposure measurements, including those by Fenske and Day (2005), OECD (1997), and U.S. EPA (1998), extensive guidance is lacking for approaches and methods for exposure assessments for pesticides in residential and other nonoccupational environments. The U.S. EPA Guidelines for Exposure Assessment (U.S. EPA, 1992) describe the general concepts of exposure assessment and provide guidance on the planning and conducting of an exposure assessment. The Guidelines focus on exposures of humans to chemical substances, but do not specifically address issues unique to particular classes of chemicals, such as pesticides. Various approaches and tools for exposure assessment and their appropriate use are discussed. It includes discussions for establishing the sampling strategy, including data quality objectives, a sampling plan, and quality assurance. The document also describes approaches for data collection, including examples of measurements to characterize various exposure-related media and parameters. The U.S. EPA Guidelines for Exposure Assessment document continues to be useful, but is out of date and does not reflect the substantial advances in exposure assessment made over the last 15plus years; but is out of date and it is currently being revised. The U.S. EPA Residential Exposure Assessment Standard Operating Procedures (SOPs) (U.S. EPA, 1997) provide guidance for assessing exposure to pesticides in a residential setting when direct measurement data are not available. The SOPs provide standard default methods for developing residential exposure assessments for both handler and postapplication exposures when chemical- and/or site-specific field data are limited. The methods in the SOPs may be used in the absence of, or as a supplement to, chemical data and/or site-specific data. Handler and postapplication SOPs for developing assessments of dermal, inhalation, and/or incidental ingestion doses are provided for major residential exposure scenarios (e.g., residential lawns, fogging, crack and crevice treatments, broadcast treatments, pet treatments, inhalation of residues from indoor treatments). A Framework for Assessing Health Risks of Environ­ mental Exposures to Children (U.S. EPA, 2006) is a useful tool in the development of the technical study design for an observational human exposure measurement study. This document discusses lifestage-specific exposure characterization and presents concepts useful for estimating children’s exposures. EPA’s Draft Protocol for Measuring Children’s NonOccupational Exposure to Pesticides by all Relevant Pathways (U.S. EPA, 2001) provides guidance for generating data for aggregate exposure assessments for children in residential environments. It provides a set of algorithms for estimating exposure by each route and pathway, describes the data needed for the estimates, and provides descriptions of approaches for data collection. It also includes examples and references for measurement methods for each exposure route (e.g., inhalation).

Chapter | 43  Sampling and Analysis for Nonoccupational Pesticide Exposure Assessments

A more recent document prepared by the U.S. EPA, Scientific and Ethical Approaches for Observational Exposure Studies (SEAOES) (U.S. EPA, 2008), describes approaches for designing and implementing observational exposure measurement studies. Although the focus of the document is on ethical issues associated with observational studies, it includes extensive discussion of the elements to be considered in study conceptualization and planning. The document stresses that the scientific and ethical approaches for these studies need to be fully integrated from the conception of the study through the final reporting and publication of results. It stresses that observational studies must be scientifically sound in order to be ethical. The document includes general information on approaches for designing and implementing observational exposure studies and citations for manuscripts and reports describing exposure measurement studies. Detailed information on alternative approaches for designing observational exposure measurement studies has not been systematically compiled and can only be obtained by a search of the scientific literature. Examples of EPA observational studies are described in the SEAOES document and also in an EPA report that summarizes results from 13 recent studies conducted or supported by EPA (U.S. EPA, 2007). During planning for the National Children’s Study (NCS), a study that will examine the impact of environmental

factors on the health of a cohort of 100,000 children from birth to 21 years of age, workgroups discussed and recommended alternative approaches for exposure measurements to address a number of different hypotheses, including hypotheses related to pesticide exposures. The results of these discussions are published in a white paper titled “Measurement and Analysis of Exposures to Environmental Pollutants and Biological Agents during the National Children’s Study” (NCS, 2004), which is available on the NCS website (http://www.nationalchildrensstudy.gov). The work of the NCS Chemical Agents workgroup was also highlighted in a mini-monograph published in Environmental Health Perspectives in 2005 (Needham et al., 2005 and others). The manuscript by Bradman and Whyatt (2005) focused on pesticides. Exposure measurement studies are complex in their design and implementation due to many factors, including differing study objectives, uniqueness of the study cohort, diversity of communities, involvement of human study participants, multiple media to be sampled for aggregate exposure estimates, and resource limitations. Figure 43.1 presents a conceptual model of residential exposures to pesticides, including elements such as mouthing activities that are important for children. The model can assist in framing the issues to be considered in design of measurement-based Respiratory tract

Inhalation

Exposure media

Source indoor residential outdoor residential industrial agricultural

Release & transfer

Inhalation rate Activity patterns

Outdoor air water soil plants surfaces

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Dermal contact Mass transfer rates

Inhalation exposure rate

R.T. uptake Uptake rate

Skin surface Dermal exposure rate

Dermal uptake Uptake rate

Absorbed dose

Contact activities Indoor air water house dust food surfaces clothes

Mouthing activities Nondietary ingestion Ingestion rate Mouthing activities

Dietary ingestion Ingestion rate Diet Exposure

Figure 43.1  Conceptual framework for children’s pesticide exposure.

Body

Ingestion rate G.I. uptake G.I. tract Ingestion exposure rate

Uptake rate

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exposure studies. Conceptualization and planning of a study involve the following elements:

Elements That May Be Included in the Study Design Introduction and background, including the purpose and scope of the study l The desired outputs and outcomes of the study, including the objectives and the hypotheses to be tested l A brief description or overview of the study l The technical approach and conceptual model that account for l sources of the chemicals being studied; l potential routes and pathways of exposure; l factors that may impact exposure and other relevant stressors; l selection and characteristics of the study participants; eligibility criteria; and recruitment, retention, and payment approaches; l justification for sample size, the methodology for selecting participants, and the sampling methods; l characteristics of the community in which the study will be performed; l environmental conditions, factors, or end points to be measured, including sampling and analysis approaches and methods (with description of expected performance); l survey design and questionnaires and other survey instruments, as applicable (with description of prior use and validation in similar studies); l pilot studies that may be undertaken; l quality assurance project plan and quality control; l time frame for the study; l exposure scenarios to be considered; l burden of the study on the participants; l resources available; and l feasibility l Discussion of alternative study designs and approaches considered and reasons for rejecting other approaches and selecting the one proposed l An analysis plan that considers l information and data needs, including data storage, security, access, and release; l nature of the measurement data (e.g., variability, quality assurance); l how the collected data will be used, and how the proposed analyses will address objectives of the study; and l hypotheses to be tested and statistical power and sample size required to test the hypotheses l Resources required or available l Project organization and management, including team members and roles and responsibilities l Schedule l

Define the study problem: Measurement studies are performed for many different purposes. During study conceptualization, the exposure science questions and problem to be addressed must be clearly identified. l Develop the study hypotheses and objectives: These are based on the science questions. l Justify the study: There must be both scientific and ethical justification for the study. To be scientifically justified, there must be a need for the study, the scientific question must not have already been answered, and the study design must be scientifically sound. Involvement of human subjects in a study must also be justified. If the hypotheses can be tested or objectives addressed without human participants in the study, then human subjects research cannot be justified. l Develop the data analysis plan: Although it may seem premature to develop a data analysis plan during the early phases of study design, the data analysis plan will determine what data should be collected. By determining how the study hypotheses will be tested or how the objectives will be addressed, the data analysis plan will identify the parameters that need to be measured (i.e., data needed) and define the required sample size(s) needed. This will in turn help to define the sampling plan. This is also referred to as identifying the “critical data elements.” l Develop the study design: The SEAOES document (U.S. EPA, 2008) describes the concepts, importance, features, and elements of a well-developed study design. Elements to be included are listed in the accompanying text box. l Prepare the human subjects research protocol, if appli­ cable: For guidance on ethical considerations in studies involving human subjects, the SEAOES document (U.S. EPA, 2008) provides extensive references for information sources. l Develop data quality objectives: Data quality objectives must be based on criteria that ensure that the data are adequate to perform the required analyses. l

The study design document should address a number of very important issues essential to the design of a scientifically sound and valid study. These include the determination of the sample size for each data element. The sample size should be sufficient to support tests for statistical significance, confidence, and other statistically based metrics. Estimates of the required sample size are important not only to ensure that the objectives of the study can be met, but also to reduce study costs by limiting the sample to only the required size (Dattalo, 2008; U.S. EPA, 2008). In addition to ensuring that the sample size is adequate to address the research objectives, the sample must be representative. Researchers must be concerned about the individuals

that participate in a study and the group or population they represent. In some cases, the study hypothesis will require a very specific study population, for example, children in child care centers with integrated pest management (IPM) practices.

Chapter | 43  Sampling and Analysis for Nonoccupational Pesticide Exposure Assessments

In other cases, the study may require that the sample be representative of the general population. Eligibility criteria for selection of study participants and the approaches and methods for recruitment must ensure both that the sample size is adequate and that the goals for representativeness are met. When the study objectives or hypotheses have been defined, the study design has been developed, and the data needs have been determined, a detailed sampling scheme is developed. The sampling scheme should systematically detail the samples to be collected, the time, location, and related sample collection logistics, and the methods to be used for sample collection. The sampling scheme should include information for both field samples and quality control samples. Quality control samples normally collected in a measurement study will include field blanks, spiked field controls, and replicate samples. If possible, performance evaluation (PE) samples prepared by an independent laboratory or standard reference materials (SRM) can also be included. Due to the nature of some of the matrices, it may not be possible to prepare spiked control samples for all sample types. A quality assurance project plan (QAPP) is required that describes the quality assurance plan for a study and describes the quality control samples and procedures to assess the accuracy and precision of the sample collection and laboratory analysis. General information on the preparation of QAPPs can be found on a number of websites, including EPA’s website http://www.epa.gov/quality/. As described previously, the draft protocol developed by EPA (U.S. EPA, 2001) for measuring children’s exposures provides guidance for identifying the parameters to be measured in order to estimate exposures by all routes and pathways. For some routes of exposure, for example inhalation, it is relatively straightforward to make estimates if the air concentrations are measured in each microenvironment, the time spent in the microenvironment is known, and there is a reasonable estimate of the individual’s inhalation rate while in the microenvironment. For other routes of exposure for example dermal, the estimates are more difficult to make and there are alternative approaches for making the estimates. The protocol presents a “macroactivity” approach and a “microactivity” approach, each of which has different data requirements. The microactivity approach, for example, requires measurements of the surface loading of the chemical, an estimate of the transfer efficiency, the surface area contacted, and the frequency of contact events. The protocol provides alternative methods for measuring these parameters. The following section discusses available methods and their application for exposure assessment.

43.3  Sample collection methods Environmental, personal, and biological samples should be selected that will account for exposure through the relevant routes and pathways based on study-specific objectives/hypotheses/scientific questions. To address the requirements of the Food Quality Protection Act of 1996,

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for example, there is a need to conduct aggregate exposure estimates in support of cumulative risk assessments. As discussed previously, aggregate exposures for pesticides may be estimated with simple algorithms or more sophisticated models that use measurements of pesticide concentrations in environmental media and diet samples. Alternatively, exposure may be estimated for some chemicals using biomonitoring data (also discussed in Chapter 45 in this handbook). In both cases, ancillary information needs to be collected to interpret the data and to make exposure estimates. Figure 43.1 depicts the potential routes of exposure (inhalation, ingestion, and dermal) and media that humans may contact. The following discussion addresses the measurements for these media, focusing primarily on the environmental media. This section describes the collection methods; the following section of the chapter (Section 43.4) describes analytical methods for identification and quantification of pesticides in the media. There are many methods available for collection of samples of environmental media, as described in reviews by Bradman and Whyatt (2005) and Lewis (2005) and in numerous manuscripts that describe measurement studies (see Tables 43.1–43.3 and the references for this chapter). It is important to note that there is little standardization of either the approaches for sample collection or the methods used to collect environmental samples for exposure assessment. Methods are generally selected based on the researcher’s familiarity with the method, available instrumentation, and, often, costs. The lack of standardization can be challenging, as it makes comparison of data sets from different studies difficult and, in many cases, may preclude conducting meta-analyses of datasets from different sources.

43.3.1  Criteria for Selection of Sample Collection Methods Numerous factors must be considered when developing a sampling plan, including sampling locations, time of collection, frequency of collection, number of samples, sample collection order, containers, potential sample contamination, and proper handling and storage of the sample prior to preparation for analysis. In addition, safety concerns, technician labor, costs, and feasibility need to be considered. Method performance needs to be fully evaluated to identify the appropriate sample collection and analysis methods. The sampling plan and selection of the methods for sample collection should be based on welldefined criteria that are likely to include the following: Study objectives: As discussed previously, the objectives determine what data are needed and what samples to collect. l Study population size: The type of instrumentation and methods used in the study will vary depending on the l

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Table 43.1  Multimedia Sample Collection Methods Matrix

Method

Key features

Reference(s)

Air

Pump with polyurethane foam (PUF) plug

Collection period up to 24 h, air volume up to 10 m3, convenient to use, less resistance to air flow

Lewis (2005)

Pump with granular sorbents: XAD, Chromosorb 102, Tenax, Porapak-R, Florisil

Collection period up to 24 h, air volume up to 10 m3, useful for collection of volatile pesticides

Lewis (2005)

Pump with PUF/ granular sorbent combination

Collection period up to 24 h, air volume up to 10 m3, extends range of compounds that can be collected

Lewis (2005)

Passive PUF sampler

PUF disk sampler deployed for integrated measurements over extended periods (weeks); has been used primarily for persistent pesticides

Jaward et al. (2004)

HVS3 (high volume surface sampler)

24-lb specialized vacuum cleaner used to collect a representative ASTM (2008d); Lewis (2005); sample of house dust. Controlled airflow and pressure drop across Roberts and Ott (2007); U.S. EPA the nozzle maintain uniform sampling conditions. High airflow (2001) volume ranges from 17 to 20 cfm. ASTM D5438

Vacuums

A number of different commercial vacuum cleaners have been adapted for controlled sample collection. Many makes and models are in use today

Roberts and Ott (2007); Lewis (2005)

Vacuum bag

For convenience and low cost, researchers collect vacuum bags from study participants. Dust concentrations can be measured, but not loading. Limited information can be collected with this method

Lewis (2005)

LWW sampler (LioyWeisel-Wainman wipe sampler)

Flat surface wipe sampler developed to measure dust on flat surfaces. Uses a template to map a specific area for the quantitative collection of dust and to control the movement of the collection plate. Sampler is constructed of Delrin with a sampling area of 109.2 cm2

Lioy et al. (1993, 2000)

EL press sampler (Edwards and Lioy press sampler)

Delrin block fitted with octadecyl (C18) extraction sheets for sample collection. Standardized 5-s press, contact pressure 0.026 lb/cm2. Small surface area of sampler limits the amount of residue collected

Edwards and Lioy (1999)

Surface wipes

Used to measure residues from hard flooring and other hard Lewis 2005; Tulve et al. (2006); residential surfaces. Many different types of wipes have been used U.S. EPA (2001) with a variety of collection protocols. Cotton gauze wetted with isopropanol has been used in a number of studies

C18 surface press sampler

Modified from the EL press sampler. Block-shaped device using C18 impregnated Teflon extraction disks as sample collection media. Surface area sampled is 114 cm2, contact pressure approximately 1200 Pa. Small surface area of sampler limits the amount of residue collected

PUF roller

Apparatus constructed of aluminum, with two permanent rear Camann et al. (1996); ASTM wheels and a detachable axle cylinder on the front where the PUF (2008c); Lewis (2005) roller is attached. Total weight of 3.9 kg for a sampling pressure of 8000 Pa. Sampling speed of 10 cm/s, used with dry sampling media, total surface area sampled is 800 cm2. ASTM D6333

House dust

Surface residues

California roller Roller is weighted with 11.4 kg for a total weight of 14.5 kg and (regular); modified and applied pressure of 2300 Pa. Sampling medium is a bed sheet laid mega samplers also on the surface. Twenty passes are made over the sampling area available

Bernard et al. (2008)

Fuller et al. (2001); Lewis (2005); Ross et al. (1991); Williams et al. (2008)

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Table 43.1  (Continued) Matrix

Method

Key features

Reference(s)

Drag sled

A block of wood or other material that is used to hold down a 10 cm  10 cm piece of denim cloth as it is dragged across the surface. A 3.6-kg weight rests on top of the block, providing a downward pressure of 4500 Pa. Sampling speed of 8 to 12 cm/s, along a 1.2-m path

Lewis (2005); Vaccaro and Cranston (1990)

Duplicate diet

Exact copy of food and liquids consumed. Solid and liquid food collected and stored in separate containers of sufficient size. Type of container dependent on contaminants of interest

Berry (1997); Bradman and Whyatt (2005); Thomas et al. (1997)

Drinking Grab sample water

An approximately 1-liter sample is collected from each unique drinking water source in a study area

Troiano et al. (2001); Bradman and Whyatt (2005); U.S. EPA (2009)

Soil

Surface scrapings

Typically, a sample collected from the top 1-cm depth in multiple locations

Simcox et al. (1995)

Dermal

Patch samplers

Considered spot or grab samples. Consists of several layers of Ferguson et al. (2007) surgical gauze and an impermeable backing (e.g., paper, cellulose, aluminum foil). Attached at specific locations on the body

Garment samplers

Range from whole-body to specific regions of the body (e.g., t-shirts, gloves, socks). Materials may be cotton, nylon, leather, blends, etc. Advantage over patch samplers is that contaminant loading over anatomical regions of the body can be conveniently collected

Wipes

Wetted wipes are moved across the body part of interest to collect Ferguson et al. (2007) contaminant from specific regions (e.g., hands, feet, knees). Little standardization of collection protocols

Rinses

One or both hands are rinsed with aqueous surfactant solutions by a variety of methods (open rinse, bag methods, wash bottles). Protocols are not standardized

Ferguson et al. (2007); Lewis (2005)

Washes

Participant washes hands according to a defined protocol and all liquid is collected and analyzed

Ferguson et al. (2007)

Fluorescent tracers

Directly and noninvasively assess dermal exposure by quantifying the deposition of fluorescent materials on the skin

Cohen Hubal et al. (2005); Ferguson et al. (2007); Lewis (2005)

Food

size of the study population and the number of samples to be collected. l Spatial variability: If concentrations of pesticide residues are expected to be highly variable across space (within a room, across rooms, for different surfaces in a room or building), it may be necessary to collect a large number of samples or to use methods that can integrate measurements across space. l Temporal variability: If long-term average concentrations are required, methods will be needed that can integrate concentrations over extended time periods. If peak concentrations or short-term fluctuations need to be measured, methods with high temporal resolution and sensitivity will be needed. l Sensitivity and detection limit: Particularly when attempting to estimate aggregate exposures and cumulative risks, method detection limits need to be

Cohen Hubal et al. (2006); Ferguson et al. (2007); Lewis (2005); Tulve et al. (2008)

sufficiently low so that the number of samples with nondetectable concentrations is minimized. This may require collection of large volumes (e.g., air samples), high mass amounts (e.g., dust), or large surface areas. l Accuracy and precision: The required performance of the method needs to be defined in data quality objectives (DQOs) based on the data analyses to be performed to address the study objectives. Related to this factor is the need for a sufficient number of appropriate quality control samples that adequately document performance of the method. l Instrument size and appropriateness for indoor moni­ toring: Methods used for personal monitoring (i.e., worn on the person) or for stationary sample collection in buildings can not be so large as to make it difficult to implement sample collection. Other factors, such as noise from pumps and sampler flow rates, must be considered for

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l



l



l



l

successful implementation in indoor environments. Security of instruments used indoors (with regard to children, pets, and curious adults) and outdoors (theft, tampering) are critical factors to consider in selection of methods. Sampler preparation and analysis requirements: Some methods require substantial preparation and labor costs. This may be related to pre-cleaning of sampling media and verification to ensure that background contamination is minimized. Some passive sampling methods, although simple to deploy and retrieve, may require extensive preparations for use. Type of sample collection method: Both passive and active sampling methods are available, primarily for air sampling. Available methods are described by Lewis (2005). Costs: Costs are a function of many variables, including the cost of equipment, sampler preparation, the analytical method, the size of the study and number of samples, and resources and labor needed to collect samples. The reality is that, in most studies, cost is a major factor in the design of the study. Burden: In the design of the sampling plan and the selection of sample collection methods, the burden on the study participant and the field team needs to be carefully evaluated. Participant burden is quantified in terms of the demands that participation in a research study places on participants with respect to their privacy, time, and efforts involved in sample collection (Dattalo, 2008). Similar to participant burden, field technician burden is used to quantify the amount of work required by the field technician to complete the study and factors into the overall costs of the study. Participants in many measurement studies are compensated for their time and efforts. Burden on participants needs to be minimized to the fullest extent to maximize the likelihood of success for a study.

43.3.2  Methods for Estimating Inhalation Exposure Methods for collection of air samples used for estimating inhalation exposures for pesticides are generally well-developed, validated, and reliable, particularly for the previous generation of semivolatile pesticides (saturation vapor pressures between 102 kPa and 108 kPa at 25°C). Air samples are collected using either active pumping or passive diffusion systems in which the air sample is collected on a sorbent media. Air samples may be collected with stationary samplers indoors or outdoors, or as a personal sample (i.e., the participant wears or carries a personal air sampler) (Lewis, 2005; U.S. EPA, 2001). As highlighted in Table 43.1, a number of different sorbent materials can be used for sample collection. XAD has been used extensively for the current-use indoor pesticides, such as the pyrethroids, which have lower vapor pressures. Sampling and analysis

methods for selected pesticides have been published as a standard practice by ASTM (2008a) and have been reviewed by Lewis (2005) and others (Table 43.1).

43.3.3  Methods for Estimating Dermal Exposure Dermal exposure monitoring techniques for assessing occupational exposure (e.g., for agricultural workers) are well-developed and used extensively (Fenske and Day, 2005; Lewis, 2005). Similar approaches have been developed and applied for estimating dermal exposure in nonoccupational environments. These include dermal patch samplers, garment samplers (covering either large portions of the body or the entire body), dermal wipes, and rinses and washes (Ferguson et al., 2007; Lewis, 2005; U.S. EPA, 2001) (Table 43.1). Patch samplers typically consist of several layers of surgical gauze and a cellulose paper backing, and represent 3–8% of the body surface area depending on the number of patches used (Ferguson et al., 2007). Researchers have used garments such as t-shirts, socks, and gloves when evaluating dermal exposure to regions of the body, and whole-body dosimeters have also been used. These garments can be made of many different materials, but the most usual are cotton, nylon, and blends (Ferguson et al., 2007; Lewis, 2005; Ross et al., 1991). Cohen Hubal et al. (2006) reported use of whole-body cotton garments to estimate young children’s potential dermal exposure as they played in child care centers. Dermal wipes, particularly hand wipes, have been used in a number of children’s studies to estimate loading of pesticide residues on the skin (Aprea et al., 1998; Fenske et al., 1986, 1998; Freeman et al., 2005; Lewis et al., 1994; Morgan et al., 2005; Wilson et al., 2007). The wipes are typically wetted with aqueous surfactant solutions (e.g., 2propanol in water) to remove the pesticide residues from the skin surface. Commercially available wipes have also been used directly without additional treatment because they are the most “nonthreatening” to very young children. The amount of residue collected by the surface wipe method will depend on the wipe used, the wetting solutions, the protocol for the method (e.g., number of wipes), and the technique of the technician doing the wiping. Therefore, it is very difficult to compare results across studies. Many different hand rinse and wash methods have also been used to estimate dermal exposure (reviewed by Lewis, 2005). The different approaches for hand rinses include “bag washes,” spray rinses with a laboratory wash bottle, open vessel rinses, and prescribed washing routines. Like the hand wipe methods, the recovery of residues with these methods is highly variable. Researchers use hand wipe, rinse, or wash data to estimate the amount of pesticide residue that a child may ingest when putting his/her hands into his/her mouth. These data may also be used to estimate the

Chapter | 43  Sampling and Analysis for Nonoccupational Pesticide Exposure Assessments

amount of pesticide residue on other parts of the body, but a number of assumptions are required for extrapolation from residues on the hand to other parts of the body. Fluorescent tracers are a noninvasive and direct means to assess dermal exposure by quantifying deposition of fluorescent materials on the skin (Cohen Hubal et al., 2005; Ferguson et al., 2007; Lewis, 2005). Fluorescent tracers are usually used in a laboratory setting where the activities of the participants can be controlled to understand how activities and surface interactions influence dermal exposure. In these laboratory studies, experiments can be designed to develop transfer coefficients (TCs) for a variety of microenvironmental/macroactivity combinations. A transfer coefficient provides a measure of dermal exposure resulting from contact with a contaminated microenvironmental surface while engaged in a specific macroactivity (U.S. EPA, 2001).

43.3.4  Indirect Methods for Estimating Dermal Exposure Estimating dermal exposure is a challenge. The direct methods of measuring residues on skin described previously are often not practical to perform in large measurement studies, particularly studies involving young children. Alternative methods that might be considered indirect methods are more easily implemented. Researchers have used different approaches for estimating dermal exposure in nonoccupational environments, including macroactivity (U.S. EPA, 2001) and microactivity (U.S. EPA, 2001; Zartarian et al., 1995, 1997) approaches. These assessment approaches provide different ways of integrating exposure over time and space. In the macroactivity approach, exposure is estimated individually for each of the microenvironments where a child spends time and each macroactivity that the child conducts within that microenvironment. To do this, exposure is modeled using empirically derived transfer coefficients to aggregate the mass transfer associated with a series of contacts with a contaminated medium. In the microactivity approach, exposure is explicitly modeled as a series of discrete transfers resulting from each contact with a contaminated medium. Both approaches require a measure of the loading of the pesticides on the surfaces being contacted. The same data on surface loadings can be used to make estimates of indirect ingestion of pesticide residues, as described in the following section. There is no standard method or approach for the measurement of surface residues. This is likely the result of the lack of satisfaction with existing methods and the failure to determine suitable methods that significantly reduce the uncertainty of the measurement and the resulting estimates of exposure. Due to the lack of standardization, comparing results across studies is difficult. Methods reported in the literature for collecting surface residues from hard surfaces (e.g., vinyl, tile, or wood), carpeted surfaces, and other surfaces (e.g., toys and

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objects children mouth) include surface wipes, press samplers, the polyurethane foam (PUF) roller, the California roller, a modified California roller, and drag sleds. These methods are highlighted in Table 43.1. When selecting the sample collection method, a number of factors need to be considered: If the surface residue measurements will be used to estimate dermal exposure by the macroactivity approach, the sample collection method should be the same as the method used to determine the empirically derived transfer coefficient (U.S. EPA, 2001). That is, if a press sampler was used to derive the transfer coefficient, a press sampler should be used to measure the surface residues. l Methods of collection need to be appropriate for the types of surfaces being monitored. Wipe methods gene­ rally are not adequate for fabric surfaces. l The recovery of pesticide residues from surfaces differs for each method and may represent different measurement parameters. The terms “total” residues and “transferrable (or dislodgeable)” residues have been used to describe the measurements (Lewis, 2005; U.S. EPA, 2001). Surface wipes, for example, may recover nearly 100% of the residues on a surface, not all of which may be available for transfer to the skin. On the other hand, methods such as the PUF roller and the press sampler were designed specifically to represent a child’s contact with a surface and the potential transfer to the skin. l The method needs to collect a representative sample, accounting for spatial variability of residues and the microenvironments where contact is most likely to occur (e.g., areas of a room where children play). l To reduce analytical costs, it may be possible, and necessary, to collect aggregate samples (e.g., with surface presses) or combine samples (e.g., surface wipes) prior to analysis to obtain “average” concentrations.

l

Detailed descriptions of methods and sampling devices for measuring surface residues can be found in Lewis (2005). The performance of surface residue sampling methods has been compared by a number of researchers, including, but not limited to, Klonne et al. (2001), Lu and Fenske (1999), and Fortune (1997) and reviewed by Lewis (2005).

43.3.5  Methods for Estimating Indirect Ingestion Indirect ingestion (also referred to as nondietary ingestion) occurs when an individual places into the mouth a hand or an object that has on its surface pesticide residues that are available for transfer to the mouth. Indirect ingestion can be estimated by determining the amount of the residue on the surface, the transfer efficiency of the residue from the object to the mouth, the area contacted, and the frequency of contacts (U.S. EPA, 2001). The wipe and rinse methods described previously can be used to determine the residue

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on the surface. The area contacted and the number of contacts have been determined in some studies by observation or with the use of video. The transfer efficiency must be determined experimentally or estimated using default “exposure factor” assumptions. Estimates of indirect ingestion may also be made using surface loadings of residues measured in dust on household surfaces and floors. Lioy et al. (2002) reviewed the importance of house dust in estimating exposures. Household dust samples have been collected by a variety of vacuum methods. The sophistication of the vacuum sampling methods ranges from use of bags from a study participant’s vacuum cleaner to collection of a dust sample from floors or furniture using the specially designed HVS3 (high volume surface sampler) or other vacuum cleaners modified specifically to collect samples under more controlled conditions (Lewis et al., 1994; Lewis, 2005; Roberts and Ott, 2007). The HVS3 is a special-purpose vacuum cleaner designed to collect house dust from surfaces in a standardized manner (ASTM, 2008c) for subsequent chemical analysis (Roberts and Ott, 2007). Lewis (2005) described a number of other handheld vacuum samplers for which the collection efficiency has been evaluated and that have been used in field measurement studies. Results of dust measurements are reported as surface loading of compound “A” (ng of A/m2) or dust concentration (ng of A/g of dust). The loading measurement is useful for estimating potential dermal exposure or indirect ingestion. The concentration measurement, although an indication of the magnitude of contamination, is not useful alone for estimating potential exposure. Table 43.1 highlights selected vacuum sample collection methods.

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observational study is being conducted. If some homes are on individual wells and others are connected to a municipal water system, then a water sample should be collected from each home with well water and one sample would be collected to represent all homes on the municipal water system. A large volume (e.g., 1 l) of water is typically collected when analyzing for pesticide residues. Water collection usually involves running the water through the system to ensure that fresh water is collected in the sample container (e.g., not water that has been standing in the pipes), collecting the water sample, preserving, if necessary, and storing at low temperature until analysis (Troiano et al., 2001). Additional guidance for collection and analysis of pesticides in drinking water (Table 43.1) is provided in ASTM (2008b) and the EPA Method 500 Series (U.S. EPA, 2009).

43.3.8  Soil Measurement Methods For the purpose of estimating exposures to pesticides in soil by the dermal, indirect ingestion, or ingestion (pica) routes of exposure, soil surface scrapings are typically collected. Core soil samples are not representative of the soil that comes into contact with the skin because of the depth at which core samples are collected. There are no standard protocols for collection of surface scrapings, although various collection protocols have been reported (Bradman and Whyatt, 2005; Lewis et al., 1994; Mukerjee et al., 1997; Simcox et al., 1995). Samples need to be collected from multiple locations to be representative of potential exposure to pesticides on outdoor soil and turf surfaces.

43.3.9  Collection of Biomonitoring Samples 43.3.6  Direct Methods for Estimating Dietary Exposure: Duplicate Diet Samples A duplicate diet sample is an exact copy of the foods and beverages that the participant eats and drinks during an observational period. The portions are identical to those consumed by the participant with respect to all aspects of preparation, type, and amount of food and drink. For sample collection purposes, it is conventional for the solid and liquid foods to be collected and stored in separate containers and for nonedible food parts to be removed before being placed in the container (e.g., bones, wrappers) as described by Thomas et al. (1997), U.S. EPA (2001), and MacIntosh et al. (2001). Berry (1997) presented an overview of the EPA’s dietary exposure program that is still currently the accepted approach for collecting duplicate diet information.

43.3.7  Water Collection Methods A water sample is typically collected from each unique water source in the geographical location where the

Biomonitoring has been used extensively to determine whether individuals have been exposed to chemicals, including pesticides (CDC, 2005). Biomarkers of exposure include measurements of pesticides, pesticide metabolites, or modified molecules or cells (e.g., protein and DNA adducts) in biological samples such as urine, blood, breath, hair, or nail clippings (Barr et al., 2005; CDC, 2005). Criteria for selection of the biomarkers and matrix to be collected are discussed by Sobus et al. in Chapter 45 of this handbook. Barr et al. (2005) discuss the various types of biomarkers that may be collected at different lifestages. For the current-use pesticides, which are generally nonpersistent and have short half-lives in the human body, pesticide metabolites are typically analyzed in urine. Methods for collection of adult urine samples are well-established and readily available. They are generally collected as spot samples or as a 24-h daily composite sample (Kissel et al., 2005), the results of which are easier to interpret. Collection of urine samples from children is more challenging. In clinical settings, urine samples can be collected with an infant urine collection bag. However, the method

Chapter | 43  Sampling and Analysis for Nonoccupational Pesticide Exposure Assessments

would be difficult to implement in a large field study with measurements collected in participants’ homes. For children who are toilet-trained, a bonnet can be inserted into the toilet for sample collection if the child is not comfortable with a direct void into a container. For younger children, cloth diapers and diapers with cotton inserts have been used (Hu et al., 2000; summarized by Barr et al., 2005). Recently, there have been advances in the development of methods to extract urine samples from the acrylate gel in disposable diapers (Hu et al., 2004).

43.3.10  Collection of Ancillary Information Such as Activity Data and Questionnaires The type of ancillary information to be collected in a measurement study is a function of the objectives of the study and the data analyses to be performed. At a minimum, to estimate exposure, it is necessary to know the concentration of the substance in the media that is contacted, the duration of contact, and the frequency of contact. The information needed to estimate the exposure depends on the route of exposure and the models that are being used to make the estimates. Using simple algorithms, inhalation exposure can be estimated using the air concentration in all occupied microenvironments, the duration of time spent in each microenvironment, and the breathing rate of the individual. Estimates of other routes of exposure, for example indirect ingestion, are more difficult and require more ancillary information on activities. Ancillary information may be collected for various purposes. Typically, questionnaires are used to collect information on sources of exposure. This information may be used in the interpretation of the study results, assessment of routes and pathways of exposure, or development or assessment of mitigation strategies or methods. Data may be collected in surveys and questionnaires to serve as surrogate metrics for parameters that cannot be measured in a study due to the complexity or the cost. For example, dietary intake may be recorded in a diary or log because costs for duplicate diet samples are high. Questionnaires are also used to collect household demographic information as well as personal information such as occupation. Although it is beyond the scope of this chapter to discuss the design and use of questionnaires and surveys, there are several good references on the topic, such as the White Paper on Measurement and Analysis of Exposures to Environmental Pollutants and Biological Agents during the National Children’s Study (NCS, 2004).

43.4  Analytical methods for pesticide measurements The analytical methods to be applied to the samples described earlier face several technical challenges because of the need

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to conduct (1) aggregate exposure assessments, addressing all routes and pathways of exposure, and (2) cumulative risk assessments that involve multiple chemicals having the same mode of toxic action. These challenges include the development of robust methods for a large suite of pesticides at or near their limits of detection, the complexity of the media/ matrices, cost limitations, wide concentration ranges, and the ever-changing suite of compounds being studied. In developing methods for ultra-trace analysis of multiple pesticide residues, the primary consideration must always be data quality. The data quality objectives for the study determine the detection limits, precision, accuracy and resolution required from the measurement. This influences everything from the selection of laboratory equipment (e.g., pipettes and balances) to the postanalysis handling of data. For these reasons, it is vital that analytical capabilities and limitations be recognized and accounted for in the study design process. The challenge arises in meeting data quality objectives for each component measured in a multi-residue suite. Though different classes of pesticides can be separated and independently processed through sample preparation procedures, there is usually enough physiochemical variation within a class to reduce the specificity of the procedure. In a clean matrix, such as solvent, this is not a concern; however, field samples may contain large quantities of other chemicals or interfering contaminants. These other chemicals may come from the environment being sampled or from the sampling media itself. The expense of collecting and analyzing samples justifies the analysis of large chemical suites, so the continuing development of improved techniques and technologies is vital. The discussion in this chapter focuses on general principles and considerations for the analyses of pesticides at trace or ultra-trace levels in the many different media collected in pesticide exposure studies. There is a wealth of information in the scientific literature describing development and application of analytical methods for pesticides. It is beyond the scope of this chapter to review and critique the many existing methods for the many different pesticide classes in environmental, food, and biological media. Existing methods are highlighted in the following sections and in Tables 43.2 and 43.3. Selected references are included to assist the reader in identifying sources of information that may be useful in evaluating and selecting methods for their specific applications.

43.4.1  General Principles Sample analysis methods typically follow the scheme shown in Figure 43.2. The first step in the analysis process is the extraction of analytes from the medium collected (e.g., air, residue, dust, food, soil). This is accomplished through solvation or by bringing in contact with a sorbent that selectively attracts the analytes of interest. Traditional methods of sample extraction, such as liquid/liquid and Soxhlet extraction, have proven to be very robust and

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Sampling media or sample matrix

Extraction

Complex extract (analytes and coextractables)

Clean-up

Clean extract (analytes and noninterfering coextractables)

Concentration

Instrument analysis

Figure 43.2  Procedure for sample analysis methods.

simple to employ, though they tend to use relatively large volumes of organic solvent and throughput is usually limited by space and glassware requirements. Current trends in sample extraction show preference to methods such as pressurized fluid extraction (PFE), solid-phase microextraction (SPME), and supercritical fluid extraction (SFE) that use substantially less solvent and are capable of much higher throughput than traditional methods (Lambropoulou and Albanis, 2007). Extract complexity is one of the most significant analytical challenges with field samples collected in observational measurement studies. In many cases, the extract from a sampling device or medium contains chemicals that change the way a target compound may perform on analytical instruments versus the way that same chemical would perform in a clean solvent. This is problematic because instruments are typically calibrated using standard solutions prepared in high-purity solvent, so any change in behavior could likely lead to erroneous quantitative results (Poole, 2007). To reduce the complexity of the sample extracts, they are purified or “cleaned-up.” Sample clean-up involves the removal of unwanted chemicals from the sample extract, thus reducing the complexity and improving the likelihood of consistent data quality. Some proven methods of cleanup include liquid partitioning, column chromatography, and gel permeation chromatography (GPC). Like the traditional extraction methods, they are very effective, but are very materials- and labor-intensive. Improved sorbent technologies and the need for higher throughput have led to the wide acceptance of solid phase extraction (SPE) as a primary clean-up technique. Most SPE methods use tubes or disks that contain a hydrophobic or hydrophilic sorbent that the analytes bind to when an extract is passed through the sorbent bed. The analytes are then eluted from the sorbent with solvent. The selection of the sorbent and solvents used to load, wash, and elute is very important and is dependent on the pesticides being analyzed and the medium from which they were extracted (Pico et al., 2007). Since the concentration of pesticides in many samples is typically very low, it is generally necessary to reduce (i.e., concentrate) the volume of sample extracts. Useful methods for sample concentration include rotary evaporation, Kuderna-Danish (K-D), nitrogen evaporation, centrifugal vacuum evaporation, and SPE. The evaporation system and conditions are an important consideration when performing pesticide analysis. Heat applied and the speed of the evaporation can have a tremendous effect on

recovery of some pesticides. One very important concentration step is fine volume adjustment, which is critical in achieving accurate volumes for clean-up or final volumes for analysis. Volumes used in SPE are very important for selectivity, and the final volume of the extract must be accurate to quantify the pesticide constituents. Sample extracts are quantitatively analyzed using techniques such as gas chromatography/mass spectrometry (GC/MS) or high-performance liquid chromatography/mass spectrometry (LC/MS). The chromatographic system separates components in an extract and the detection system provides a measurement of each component. The high degree of selectivity along with the ability to confirm chemical identity make mass spectrometers the detection system of choice since concentrations that occur in nonoccupational exposure samples are usually very low. The separation process is very important since many pesticides within a class can exhibit similar characteristics that would make them indistinguishable, even when using mass spectrometers. GC/MS is the preferred means of analysis because of unmatched separation efficiency, high reliability and minimal maintenance, mass spectrometric simplicity from analytes since they are already in the gas phase, and higher degree of automation because the mobile phase (a compressed gas) does not have to be prepared and can last for months from a single cylinder. GC/MS is a very powerful analytical technique but also has some key limitations. In order to be analyzed by GC, a compound must be volatile, thermally stable, and relatively nonpolar. For many years, these limitations were not a great concern in pesticide analysis, since most pesticides were amenable to analysis by GC methods. The current trend toward LC/MS analysis arose from the need to measure metabolites of nonpersistent pesticides in biological media (Hernandez et al., 2005). Most pesticide metabolites are nonvolatile and polar (Soler et al., 2008), so without derivatization, they cannot be analyzed by GC. Other factors that have led to the increasing popularity of LC/MS are improvements in interfaces, chromatographic columns, and mass analyzer technologies. These improvements have made LC/MS as robust as it is versatile, so applications that would have typically been performed by GC/MS are being consolidated into LC/MS methods. A practical application of this is the combination of parent and metabolite analysis from the same sample in a single analytical run (Jansson et al., 2004). The analysis of both in the environmental and biological sample can help to better correlate the two measurements.

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Table 43.2  Methods and Method Reviews for Measuring Pesticides in Exposure Studies Matrix

Analytes

Method description

Analysis method

Reference

Food

Organophosphates (OPs) DDT

PFE, ASE

GC/MS ELISA

Chuang et al. (2001)

OPs, pyrethroids, triazoles, triazines

QueChERS

GC/MS

Hercegova et al. (2006)

Multiclass pesticides, PCBs

LLE, Soxhlet, PFE, SFE, GPC (review)

Not specified

Beyer and Biziuk (2008)

Multiclass pesticides

SPE (review)

GC/MS, GC/ECD, LC/MS, CE

Pico et al. (2007)

Multiclass pesticides

Detailed LC/MS (review)

LC/MS

Soler et al. (2008)

Multiclass pesticides

Soxhlet, QuEChERS, SFE, ultrasonication, GPC, SPE (review)

GC/MS, LC/MS

Lambropoulou and Albanis (2007)

Multiclass pesticides

ELISA specific – ranges and MDLs (review)

ELISA

Morozova et al. (2005)

Multiclass pesticide residues

SPE, column switching, on-line SPE (review)

LC/MS/MS

Pico et al. (2004)

Multiclass pesticides

CE specific (review)

CE

Malik and Faubel (2001)

Organophosphates, PCBs

Extraction, clean-up, QA considerations (review)

GC/MS, GC/ECD, ELISA

Muir and Sverko (2006)

Multiclass pesticides

Sample description and collection considerations (review)

Not specified

Bradman and Whyatt (2005)

Organophosphates, pyrethroids

Sampling and extraction, PFE

GC/ECD

Bernard et al. (2008)

Chlorpyrifos

Soxhlet, shake-flask extraction

GC/ECD

Stout and Mason (2003)

Water

Carbamates and carbamate metabolites

SPE extraction (review)

LC/MS: interfaces discussed

Soriano et al. (2001)

Dust

Pyrethroids and pyrethroid metabolites

Extraction, clean-up, derivitization

GC/MS

Starr et al. (2008)

Nonspecific

Multiclass pesticides

LC/MS and GC/MS comparison by pesticide (review)

GC/MS, LC/MS LODs and LOQs for many pesticides

Alder et al. (2006)

Multimedia (review)

ASE, accelerated solvent extraction; CE, capillary electrophoresis; ELISA, enzyme-linked immunosorbant assay; GC/ECD, gas chromatography/electron capture detector; GC/MS, gas chromatography/mass spectrometry; GPC, gel permeation chromatography; LC/MS, liquid chromatography/mass spectrometry; LC/MS/MS, liquid chromatography/tandem mass spectrometry; LLE, liquid–liquid extraction; LODs, limits of detection; LOQs, limits of quantitation; MDLs, method detection limits; PFE, pressurized fluid extraction; QueChERS, quick, easy, cheap, effective, rugged, and safe; SFE, supercritical fluid extraction; SPE, solid phase extraction.

Table 43.2 highlights selected analytical methods for measurements of pesticides in food and environmental media. Table 43.3 presents methods published for biological media.

43.4.2  Method Performance Requirements As noted previously, the measurement methods, which include both the sample collection and the analytical methods, must meet well-defined data quality requirements in

order to address study objectives. Data quality objectives need to be defined for accuracy, precision, and completeness. Additionally, limits of quantitation and detection limits need to be determined. To meet data analysis objectives, the limitations of both the sample collection and sample analysis methods need to be identified. In many cases, the sample collection method may be the limiting factor in meeting data quality objectives, particularly with regard to detection limits (e.g., collection of adequate mass of floor dust for analysis).

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Table 43.3  Methods and Method Reviews for Analysis in Biological Media Matrix

Analytes

Method description

Analysis method

Reference

Multiple biological

Multiclass pesticides and metabolites

Extraction, clean-up, and analysis

GC/ECD, GC/MS, HPLC/UV, LC/MS

Aprea et al. (2002)

Multiclass pesticides and metabolites

Emphasis on LC/MS

LC/MS

Hernandez et al. (2005)

Multiclass pesticides and metabolites

Sample extraction, preparation, and analysis considerations by class

GC/MS,GC/FPD, GC/NPD, GC/ECD, LC/MS

Barr and Needham (2002)

Multiclass pesticides and metabolites

Discussion by class

GC/MS, LC/MS, HPLC/UV, HPLC/EC

Barr (2008)

Organophosphates

SPE, derivitization

GC/MS/MS

Hemakanthi De Alwis et al. (2006)

Multiclass metabolites

Preparation and analysis, labeled analogs

GC/MS/MS

Shealy et al. (1996)

Cis- CDCA, transCDCA

Ultrasonic extraction, SPE, derivitization

GC/MS-NCI

Elflein et al. (2003)

Organochlorine pesticides and other halogenated pollutants

PFE, GPC, SPE

GC/MS

Saito et al. (2004)

Urine

Adipose, heart, kidney, liver

GC/ECD, gas chromatography/electron capture detector; GC/FPD, gas chromatography/flame photometric detector; GC/MS, gas chromatography/mass spectrometry; GC/MS/MS, gas chromatography/tandem mass spectrometry; GC/MS-NCI, gas chromatography/mass spectrometry—negative chemical ionization; GC/NPD, gas chromatography/nitrogen phosphorous detector; GPC, gel permeation chromatography; HPLC/EC, high-performance liquid chromatography/electrochemical detector; HPLC/UV, high-performance liquid chromatography/ultraviolet detector; LC/MS, liquid chromatography/ mass spectrometry;PFE, pressurized fluid extraction; SPE, solid phase extraction.

Aside from the technical considerations regarding analyte detection, there has been a copious amount of discussion regarding the definition of detection limit and data reporting at or below the detection limit. The U.S. EPA method detection limit (MDL) procedure can be found in Title 40 Code of Federal Regulations (40 CFR 136, Appendix B, revision 1.11). Despite some criticism of this procedure, it remains a simple, well-documented way to determine method detection limits. Another significant issue for consideration is data censoring resulting from nondetects (Helsel, 1990). Although this does not affect the handling of data in the analytical laboratory, it significantly impacts analyses of data from aggregate and cumulative exposure studies employing multi-residue analyses in multiple media where there can be a potentially large number of samples with concentrations below the method detection limit. This issue indicates the continuing need for development of methods with lower detection limits. Analytical methods are needed that have the capability to measure multiple pesticide residues in a single sample. Requirements for performing cumulative risk assessments and measuring exposures to multiple pesticides with the same toxicological endpoint, in addition to the high level of effort and the high costs associated with sample collection and analysis, make this a necessity. The analytical performance for a suite of pesticides needs to be evaluated early

in the design of a measurement program. Performance may vary for different pesticides in the suite, necessitating decisions on what measurements to perform, the selection of sample collection methods, and determination of whether the data analysis objectives can be met for all pesticides proposed for the measurement study. In addition to achieving acceptable detection limits, these analytical methods must be developed for a variety of media. Though some methods can be used for a variety of matrices, specialized methods are typically developed for each type of medium. This is done because of the wide variation in the physical qualities of media (solid/liquid/gas) as well as the materials that are co-extracted or co-soluble with targeted pesticides in each. A key consideration in selection of sampling and analysis methods is the ability to obtain sampling media that are of consistent quality and have minimal background contamination and/or interferences with the analytical method. Intermedia variations in purchased media such as surface wipes and PUF need be accounted for. Standardization is highly recommended due to batch-to-batch variation in sampling media as well as background contamination. Standardization includes cleaning and storage of media in the same way prior to field deployment. Media may be purchased that is pre-cleaned or batches may be prepared in the analytical lab. It is critical that samples from all

Chapter | 43  Sampling and Analysis for Nonoccupational Pesticide Exposure Assessments

batches of media be analyzed for the target analytes prior to use to confirm that there is no, or at least minimal, background contamination.

43.4.3  State-of-the-Science Current analytical trends reflect the need to measure nonpersistent pesticides and their degradation products. This is evidenced by the increasing number of methods being developed for application to biological matrices and development of methods for analyses of both the parent compounds and metabolites/degradation products in biological and environmental samples. Table 43.3 highlights methods for biological media and provides selected literature references, including review articles. Advances in biological monitoring methods have been significant in recent years due to improvements in LC/MS interfaces and the commercial availability of reference standards. Advances in analytical technology have made it easier to deal with the ever-changing suite of analytes being studied. New extraction, sample preparation, and analytical systems are capable of automating tedious tasks and improving throughput. This can translate into more efficient methods of development and can eliminate human bias in many processes. There are certain limitations to these new systems, primarily due to the size of the sampling media, which are usually large (e.g. cotton garments, PUF), and the content of co-extractable chemicals. This is an important consideration since the acquisition cost of many of these new instruments is significant. In addition to better availability of analytical reference standards for pesticides, stable isotopes increasingly are becoming commercially available. Although they can be very expensive, isotopically labeled analogs of pesticides can be invaluable when used as internal standards or surrogate recovery standards. Because they are not naturally occurring, the labeled standards can be used to normalize responses or quantify recoveries without the concern of interference. Because they behave in the same way as the unlabeled pesticide, anything that happens to the pesticide during the analytical method will also happen to the standard. This allows for the identification, and in many cases, correction of errors in sample preparation or instrumental analysis. The availability of these labeled analogs is another factor that has made mass spectrometry the primary detection method used. Two-dimensional detection systems cannot distinguish between a pesticide and its labeled analog if they co-elute, which is considered the ideal situation to compensate for instrument effects. Much research into immunochemical methods has occurred in the last few years. Enzyme-linked immunosorbant assay (ELISA), in particular, has advanced to a point where many commercially available kits can be purchased for pesticide analysis. The advantages of ELISA are very low detection limits and high selectivity. The selectivity is a disadvantage when applied to multi-residue methods

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since a suite of chemicals cannot be measured in a single test. Other disadvantages are the limited analytical range and quantitative limitations due to matrix effects. Immunochemical methods and technologies are improving and may become more viable alternatives in the future (Van Emon, 2006).

Conclusion Direct measurements of pesticides in environmental media and diet continue to be an important tool for estimating human exposure and for determining the factors that have the greatest impact on people’s exposures. Unlike worker exposure, for which there are several reviews and guidance documents on methodology for exposure measurements, extensive guidance is lacking for approaches and methods for exposure assessments for pesticides in residential and other nonoccupational environments. The need to estimate aggregate and cumulative exposures to pesticides challenges researchers to develop and validate systematic sample collection protocols, as well as new sampling and analysis methods. Some guidance on sampling protocols has been published, but in general, neither the sample collection protocols nor the collection methods have been standardized. As a result comparison of results collected by different researchers is a challenge. The sample collection methods highlighted in this chapter are not new to the field of exposure assessment (e.g., these methods have been reported in the literature dating back to the early 1990s), and all have been described at great length in various review articles (including Bradman and Whyatt, 2005; Lewis, 2005). While many of the methods appear adequate, there is a need to advance and standardize sample collection methods. For example, surface wipes are routinely used for pesticide residue sample collection, but there are multiple methods in use and the collection protocols are not standardized. There have been few reports of systematic evaluation of the wipes, documenting method performance in terms of accuracy and precision. Similarly, methods and protocols for hand wipes and rinses are highly variable. There have been a number of reports documenting the performance of vacuum dust collection methods, but many of these were for previous-generation pesticides (e.g., the organophosphates). The data for collection efficiency of these methods for current residentialuse pesticides, such as the pyrethroids, are limited. In spite of the large number of measurement studies in which pesticides have been measured in various media during the past decade, it appears that advances in sample collection methods have been limited. Many of the methods still require substantial investment in equipment, materials, and labor to implement. Advances are needed for lower cost and lower burden methods that can be used in larger measurement studies, such as the National Children’s Study.

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Significant advances have been made in development and refinement of analytical methods. Multi-residue methods have been published by a number of researchers. These methods have been adapted for many matrices. Improved sensitivity of the methods has been reported. There have also been significant advances in the analysis of pesticide degradation products (metabolites) in biological media, environmental media, and diet samples. With the expansion of biomonitoring programs and the desire to use biomarkers to estimate exposures, it is increasingly important that the metabolite concentrations also be quantified in environmental media and diet in order to interpret biomonitoring results. Sampling and analysis of pesticides and their degradation products will continue to provide the critical information needed to protect public health and the environment. However, development of sampling and analysis methods for pesticides will continue to be a challenge with the nearly continuous introduction of new pesticide active ingredients and formulations.

Acknowledgments This paper has been reviewed in accordance with the United States Environmental Protection Agency’s Office of Research and Development peer and administrative review policies and approved for publication.

References Alder, L., Greulich, K., Kempe, G., and Vieth, B. (2006). Residue analysis of 500 high priority pesticides: Better by GC-MS or LC-MS/MS. Mass Spect. Rev. 25, 838–865. Aprea, C., Colosio, C., Mammone, T., Minoia, C., and Maroni, M. (2002). Biological monitoring of pesticide exposure: A review of analytical methods. J. Chromatogr. B 769, 191–219. Aprea, C., Sciarra, G., Sartorelli, P., Mancini, R., and Di Luca, V. (1998). Environmental and biological monitoring for mancozeb, ethylenethiourea and dimethoate during industrial formulation. J. Toxicol. Environ. Health 53, 263–281. ASTM. (2008a). Standard Practice for Selection of Analytical Techniques for Pesticides and Polychlorinated Biphenyls in Air. ASTM D4861. In “Annual Book of ASTM Standards.” Vol. 11.03. ASTM, West Conshohocken, PA. ASTM. (2008b). Standard Test Method for Determination of Organochlorine Pesticides in Water by Capillary Column Gas Chromatography. ASTM D5812. In “Annual Book of ASTM Standards.” ASTM, West Conshohocken, PA. ASTM. (2008c). Standard Practice for Collection of Dislodgeable Residues from Floors. ASTM D6333. In “Annual Book of ASTM Standards.” Vol. 11.03. ASTM, West Conshohocken, PA. ASTM. (2008d). Standard Practice for Collection of Floor Dust for Chemical Analysis. ASTM D5348. In “Annual Book of ASTM Standards.” Vol. 11.03. ASTM, West Conshohocken, PA. Barr, D. B. (2008). Biomonitoring of exposure to pesticides. J. Chem. Health Safety (November/December), 20–29.

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Barr, D. B., Wang, R. Y., and Needham, L. L. (2005). Biologic monitoring of exposure to environmental chemicals throughout the life stages: requirements and issues for consideration for the national children’s study. Environ. Health Perspect. 113(8), 1083–1091. Barr, D., and Needham, L. (2002). Analytical methods for biological monitoring of exposure to pesticides: A review. J. Chromatogr. B 778, 5–29. Bernard, C. E., Berry, M. R., Wymer, L. J., and Melnyk, L. J. (2008). Sampling household surfaces for pesticide residues: comparison between a press sampler and solvent-moistened wipes. Sci. Total Environ. 389(2–3), 514–521. Berry, M. R. (1997). Advances in dietary exposure research at the United States Environmental Protection Agency-National Exposure Research Laboratory. J. Expo. Anal. Environ. Epidemiol. 7(1), 3–16. Beyer, A., and Biziuk, M. (2008). Applications of sample preparation techniques in the analysis of pesticides and PCBs in food. Food Chem. 108, 669–680. Bradman, A., and Whyatt, R. M. (2005). Characterizing exposures to nonpersistent pesticides during pregnancy and early childhood in the National Children’s Study: a review of monitoring and measurement methodologies. Environ. Health Perspect. 113(8), 1092–1099. Camann, D.E., Harding, H.J., Geno, P.W., Agrawal, S.R. (1996). “Comparison of Methods to Determine Dislodgeable Residue Transfer from Floors.” Research Triangle Park, NC, US Environmental Protection Agency. EPA/600/R-96/089. CDC. (2005). “Third National Report on Human Exposure to Environmental Chemicals.” Atlanta, Georgia: National Center for Environmental Health, Division of Laboratory Sciences. NCEH Pub. No. 05-0570. Chuang, J. C., Hart, K., Chang, J. S., Boman, L. E., Emon, J. M. V., and Reed, A. W. (2001). Evaluation of analytical methods for determining pesticides in baby foods and adult duplicate diet samples. Anal. Chim. Acta 444, 87–95. Cohen Hubal, E. A., Egeghy, P. P., Leovic, K. W., and Akland, G. G. (2006). Measuring potential dermal transfer of a pesticide to children in a child care center. Environ. Health Perspect. 114(2), 264–269. Cohen Hubal, E. A., Suggs, J. C., Nishioka, M. G., and Ivancic, W. A. (2005). Characterizing residue transfer efficiencies using a fluorescent imaging technique. J. Expo. Anal. Environ. Epidemiol. 15(3), 261–270. Dattalo, P. (2008). “Determining Sample Size: Balancing Power, Precision & Practicality.” Oxford University Press, U.S. Edwards, R. D., and Lioy, P. J. (1999). The EL sampler: a press sampler for the quantitative estimation of dermal exposure to pesticides in housedust. J. Expo. Anal. Environ. Epidemiol. 9(5), 521–529. Elflein, L., Berger-Preiss, E., Preiss, A., Elend, M., Levsen, K., and Wunsch, G. (2003). Human biomonitoring of pyrethrum and pyrethroid insecticides used indoors: Determination of metabolites E-cis/trans-chrysanthemumdicarboxylic acid in human urine by gas chromatography-mass spectrometry with negative chemical ionization. J. Chromatogr. B 795, 195–207. Fenske, R. A., and Day, E. W. Jr., (2005). Assessment of exposure for pesticide handlers in agricultural, residential, and institutional environments Chapter 1. In “Occupational and Residential Exposure Assessment for Pesticides” (C. A. Franklin and J. P. Worgan, eds.), pp. 13–43. John Wiley & Sons, Inc., England. Fenske, R. A., Schulter, C., Lu, C., and Allen, E. H. (1998). Incomplete removal of the pesticide captan from skin by standard handwash exposure assessment procedures. Bull. Environ. Contam. Toxicol. 61, 194–201. Fenske, R. A., Wong, S. M., Leffingwell, J. T., and Spear, R. C. (1986). A video imaging technique for assessing dermal exposure – II. Fluorescent tracer testing. Am. Ind. Hyg. Assoc. J. 47, 771–775.

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Ferguson, A. C., Canales, R. A., and Leckie, J. O. (2007). Dermal exposure, uptake, and dose, Chapter 11. In “Exposure Analysis” (ε. R. Ott, A. C. Steinemann, and L. A. Wallace, eds.), pp. 255–284. Florida: Taylor & Francis Group, CRC Press. Fortune, C. R. (1997). “Evaluation of Methods for Collecting Dislodgeable Pesticide Residues from Turf.” Research Triangle Park, NC: US Environmental Protection Agency. EPA/600/R-97/108. Freeman, N. C., Hore, P., Black, K., Jimenez, M., Sheldon, L., Tulve, N., and Lioy, P. J. (2005). Contributions of children’s activities to pesticide hand loadings following residential pesticide application. J. Expo. Anal. Environ. Epidemiol. 15(1), 81–88. Fuller, R., Klonne, D., Rosenheck, L., Eberhart, D., Worgan, J., and Ross, J. (2001). Modified California roller for measuring transferable residues on treated turfgrass. Bull. Environ. Contam. Toxicol. 67(6), 787–794. Helsel, D. R. (1990). Less than obvious-statistical treatment of data below the detection limit. Environ. Sci. Technol. 24(12), 1766–1774. Hemakanthi De Alwis, G. K., Needham, L., and Barr, D. (2006). Measurement of human urinary organophosphate pesticide metabolites by automated solid-phase extraction, post extraction derivitization, and gas chromatography-tandem mass spectrometry. J. Chromatog. B 843, 34–41. Hercegova, A., Domotorova, M., Kruzlicaova, D., and Matisova, E. (2006). Comparison of sample preparation methods combined with fast gas chromatography-mass spectrometry for ultratrace analysis of pesticide residues in baby food. J. Sep. Sci. 29, 1102–1109. Hernandez, F., Sancho, J. V., and Pozo, O. J. (2005). Critical review of the application of liquid chromatography/mass spectrometry to the determination of pesticide residues in biological samples. Anal. Bioanal. Chem. 382, 934–946. Hu, Y. A., Barr, D. B., Akland, G., Melnyk, L., Needham, L., Pellizzari, E. D., Raymer, J. H., and Roberds, J. M. (2000). Collecting urine samples from young children using cotton gauze for pesticide studies. J. Expo. Anal. Environ. Epidemiol. 10(6 Pt 2), 703–709. Hu, Y., Beach, J., Raymer, J., and Gardner, M. (2004). Disposable diaper to collect urine samples from young children for pyrethroid pesticide studies. J. Expo. Anal. Environ. Epidemiol. 14(5), 378–384. Jansson, C., Pihlstrom, T., Osterdahl, B.-G., and Markides, K. E. (2004). A new multi-residue method for analysis of pesticide residues in fruit and vegetables using liquid chromatography with tandem mass spectrometric detection. J. Chromatogr. A 1023, 93–104. Jaward, F. M., Farrar, N. J., Harner, KT., Sweetman, A. J., and Jones, K. C. (2004). Passive air sampling of PCBs, PBDEs, and organochlorine pesticides across Europe. Environ. Sci. Technol. 38(1), 34–41. Kissel, J. C., Curl, C. L., Kedan, G., Lu, C., Griffith, W., Barr, D. B., Needham, L. L., and Fenske, R. A. (2005). Comparison of organophosphorus pesticide metabolite levels in single and multiple daily urine samples collected from preschool children in Washington State. J. Expo. Anal. Environ. Epidemiol. 15(2), 164–171. Klonne, D., Cowell, J., Mueth, M., Eberhart, D., Rosenheck, L., Ross, J., and Worgan, J. (2001). Comparative study of five transferable turf residue methods. Bull. Environ. Contam. Toxicol. 67(6), 771–779. Lambropoulou, D., and Albanis, T. (2007). Methods of sample preparation for determination of pesticide residues in food matrices by chromatography-mass spectrometry-based techniques: A review. Anal. Bioanal. Chem. 389, 1663–1683. Lewis, R. J., Fortmann, R. C., and Camann, D. E. (1994). Evaluation of methods for monitoring the potential exposure of small children to pesticides in the residential environment. Environ. Contam. Toxicol. 26, 37–46. Lewis, R. G. (2005). Residential post-application pesticide exposure monitoring, Chapter 3. In “Occupational and Residential Exposure

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Assessment for Pesticides” (C. A. Franklin and J. P. Worgan, eds.), pp. 71–128. John Wiley & Sons, Inc., England. Lioy, P. J., Edwards, R. D., Freeman, N., Gurunathan, S., Pellizzari, E., Adgate, J. L., Quackenboss, J., and Sexton, K. (2000). House dust levels of selected insecticides and a herbicide measured by the EL and LWW samplers and comparisons to hand rinses and urine meta­ bolites. J. Expo. Anal. Environ. Epidemiol. 10(4), 327–340. Lioy, P. J., Freeman, N. C. G., and Millette, J. R. (2002). Dust: a metric for use in residential and building exposure assessment and source characterization. Environ. Health Perspect. 110(10), 969–983. Lioy, P. J., Wainman, T., and Weisel, C. (1993). A wipe sampler for the quantitative measurement of dust on smooth surfaces: laboratory performance studies. J. Expo. Anal. Environ. Epidemiol. 3(3), 315–330. Lu, C., and Fenske, R. A. (1999). Dermal transfer of chlorpyrifos residues from residential surfaces: comparison of hand press, hand drag, wipe, and polyurethane foam roller measurements after broadcast and aerosol pesticide applications. Environ. Health Perspect. 107(6), 463–467. MacIntosh, D. L., Kabiru, C. W., and Ryan, P. B. (2001). Longitudinal investigation of dietary exposure to selected pesticides. Environ. Health Perspect. 109(2), A85. Malik, A., and Faubel, W. (2001). A review of analysis of pesticides using capillary electrophoresis. Crit. Rev. Anal. Chem. 31(3), 223–279. Morgan, M. K., Sheldon, L. S., Croghan, C. W., Jones, P. A., Robertson, G. L., Chuang, J. C., Wilson, N. K., and Lyu, C. W. (2005). Exposures of preschool children to chlorpyrifos and its degradation product 3,5,6-trichloro-2-2pyridinol in their everyday environments. J. Expo. Anal. Environ. Epidemiol. 15(4), 297–309. Morozova, V. S., Levashova, A. I., and Eremin, S. A. (2005). Determination of pesticides by enzyme immunoassay. J. Anal. Chem. 60(3), 202–217. Muir, D., and Sverko, E. (2006). Analytical methods for PCBs and organochlorine pesticides in environmental monitoring and surveillance: A critical appraisal. Anal. Bioanal. Chem. 386, 769–789. Mukerjee, S., Ellenson, W. D., Lewis, R. G., Stevens, R. K., Somerville, M. C., Shadwick, D. S., and Willis, R. D. (1997). An environmental scoping study in the lower Rio Grande Valley of Texas – III. Residential microenvironmental monitoring for air, house dust, and soil. Environ. Int. 23(5), 657–673. NCS. (2004). Final White Paper with Executive Summary: Measurement and Analysis of Exposures to Environmental Pollutants and Biological Agents during the National Children’s Study (November 2004), available at http://www.nationalchildrensstudy.gov/research/reviewsreports/ Pages/default.aspx. Needham, L. L., Ozkaynak, H., Whyatt, R. M., Barr, D. B., Wang, R. Y., Naeher, L., Akland, G., Bahadori, T., Bradman, A., Fortmann, R., Liu, L. J. S., Morandi, M., O’Rourke, M. K., Thomas, K., Quackenboss, J., Ryan, P. B., and Zartarian, V. (2005). Exposure Assessment in the National Children’s Study: Introduction. Environ. Health Perspect. 113(8), 1076–1082. OECD. (1997). “Guidance Document for the Conduct of Occupational Exposure to Pesticides During Agricultural Application,” OECD Environmental Health and Safety Publications, Series on Testing and Assessment, No. 9, OCDE/GD(77)148, Paris, France. Pico, Y., Biasco, C., and Font, G. (2004). Environmental and food applications of LC-tandem mass spectrometry in pesticide-residue analysis: An overview. Mass Spectrom. Rev. 23, 45–85. Pico, Y., Fernandez, M., Ruiz, M. J., and Font, G. (2007). Current trends in solid-phase-based extraction techniques for the determination of pesticides in food and environment. J. Biochem. Biophys. Methods 70, 117–131.

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Poole, C. F. (2007). Matrix-induced response enhancement in pesticide residue analysis by gas chromatography. J. Chromatogr. A. 1158, 241–250. Roberts, J. W., and Ott, W. R. (2007). Exposure to pollutants from house dust, Chapter 14. In “Exposure Analysis” (W. R. Ott, A. C. Steinemann, and L. A. Wallace, eds.), pp. 319–345. Taylor & Francis Group, CRC Press, Florida. Ross, J., Fong, H. R., Thongsinthusak, T., Margetich, S., and Krieger, R. I. (1991). Measured potential dermal transfer of surface pesticide residue generated from indoor fogger use: using the CDFA roller method. Chemosphere 22, 975–984. Saito, K., Sjodin, A., Sandau, C. D., Davis, M. D., Nakazawa, H., Matsuki, Y., and Patterson, D. G. Jr. (2004). Development of a accelerated solvent extraction and gel permeation chromatography analytical method for measuring persistent organohalogen compounds in adipose and organ tissue analysis. Chemosphere 57, 373–381. Shealy, D. B., Bonin, M. A., Wooten, J. V., Ashley, D. L., Needham, L. L., and Bond, A. E. (1996). Application of an improved method for the analysis of pesticides and their metabolites in the urine of farmer applicators and their families. Environ. Int. 22(6), 661–675. Simcox, N. J., Fenske, R. A., Wolz, S. A., Lee, I., and Kalam, D. A. (1995). Pesticides in household dust and soil: exposure pathways for children of agricultural families. Environ. Health Perspect. 103(12), 1126–1134. Soler, C., Manes, J., and Pico, Y. (2008). The role of liquid chromatographymass spectrometry in pesticide residue determination in food. Crit. Rev. Anal. Chem. 38, 93–117. Soriano, J., Jiminez, B., Font, G., and Molto, J. (2001). Analysis of carbamate pesticides and their metabolites in water by solid phase extraction and liquid chromatography: A review. Crit. Rev. Anal. Chem. 31, 19–32. Starr, J., Graham, S., Stout, D. II, Andrews, K., and Nishioka, M. (2008). Pyrethroid pesticides and their metabolites in vacuum cleaner dust collected from homes and day-care centers. Environ. Res. 108, 271–279. Stout, D. M., and Mason, M. A. (2003). The distribution of chlorpyrifos following a crack and crevice type application in the US EPA Indoor Air Quality Research House. Atmos. Environ. 37, 5539–5549. Thomas, K. W., Sheldon, L. S., Pellizzari, E. D., Handy, R. W., Roberds, J. M., and Berry, M. R. (1997). Testing duplicate diet sample collection methods for measuring personal dietary exposures to chemical contaminants. J. Expo. Anal. Environ. Epidemiol. 7(1), 17–36. Troiano, J., Weaver, D., Marade, J., Spurlock, F., Pepple, M., Nordmark, C., and Bartkowiak, D. (2001). Summary of well water sampling in California to detect pesticide residues resulting from nonpoint-source applications. J. Environ. Qual. 30, 448–459. Tulve, N. S., Egeghy, P. P., Fortmann, R. C., Whitaker, D. A., Nishioka, M. G., Naeher, L. P., and Hilliard, A. (2008). Multimedia measurements and activity patterns in an observational pilot study of nine young children. J. Expo. Sci. Environ. Epidemiol. 18(1), 31–44. Tulve, N. S., Jones, P. A., Nishioka, M. G., Fortmann, R. C., Croghan, C. W., Zhou, J. Y., Fraser, A., Cave, C., and Friedman, W. (2006). Pesticide measurements from the first national environmental health survey of child care centers using a multi-residue GC/MS analysis method. Environ. Sci. Technol. 40(20), 6269–6274.

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U.S. EPA. (1992). Guidelines for Exposure Assessment. Risk Assessment Forum, U.S. Environmental Protection Agency, Washington, DC. EPA/600/Z-92/001, May 1992. U.S. EPA. (1997). Standard Operating Procedures (SOPs) for Residential Exposure Assessments. Office of Prevention, Pesticides, and Toxic Substances, U.S. Environmental Protection Agency, Washington, DC. http://www.epa.gov/scipoly/sap/meetings/1997/september/sopindex.htm. U.S. EPA. (1998). Post Application Guidelines: Series 875 – Group B – Occupational and Residential Exposure Test Guidelines, Version 5.4, Office of Prevention, Pesticides, and Toxic Substances, Washington, DC. U.S. EPA. (2001). Draft Protocol for Measuring Children’s NonOccupational Exposure to Pesticides by all Relevant Pathways. Research Triangle Park, NC: Office of Research and Development. EPA/600/R-03/026. U.S. EPA. (2006). A Framework for Assessing Health Risks of Environmental Exposure to Children. Washington, DC: US EPA. EPA/600/R-05/093F. http://cfpub.epa.gov/ncea/cfm/recordisplay. cfm?deid158363. U.S. EPA. (2007a). Important Exposure Factors for Children. An Analysis of Laboratory and Observational Field Data Characterizing Cumulative Exposure to Pesticides. Washington, DC: Office of Research and Development. EPA/600/R-07/013. U.S. EPA. (2008). The Scientific and Ethical Approaches for Observational Exposure Studies. EPA Report No. EPA/600/R-08/062. U.S. Environmental Protection Agency, Office of Research and Development, Research Triangle Park, NC. http://www.epa.gov/nerl/sots. U.S. EPA. (2009). Safe Drinking Water Analytical Methods and Laboratory Certification, U.S. Environmental Protection Agency website: http://www.epa.gov/safewater/methods/index.html. Vaccaro, J. R., and Cranston, R. J. (1990). “Evaluation of dislodgeable residues and absorbed doses of chlorpyrifos following indoor broadcast applications of chlorpyrifos-based emulsifiable concentrate.” Dow Chemical Company, Midland, MI. Van Emon, J. M. (ed.) (2006). “Immunoassay and Other Bioanalytical Techniques.” CRC Press, Boca Raton, FL. Williams, R. L., Bernard, C. E., Dyk, M. B., Ross, J. H., and Krieger, R. I. (2008). Measurement of transferable chemical residue from nylon carpet using the California roller and a new mega-California roller. J. Environ. Sci. Health B. 43(8), 675–679. Wilson, N. K., Chuang, J. C., Morgan, M. K., Lordo, R. A., and Sheldon, L. S. (2007). An observational study of the potential exposures of preschool children to pentachlorophenol, bisphenol-A, and nonylphenol at home and daycare. Environ. Res. 103(1), 9–20. Zartarian, V. G., Ferguson, A. C., Ong, C. G., and Leckie, J. O. (1997). Quantifying videotaped activity patterns: video translation software and training methodologies. J. Expo. Anal. Environ. Epidemiol. 7(4), 535–542. Zartarian, V. G., Streicker, J., Rivera, A., Cornejo, C. S., Molina, S., Valadez, O. F., and Leckie, J. O. (1995). A pilot study to collect micro-activity data of two- to four-year-old farm labor children in Salinas Valley, California. J. Expo. Anal. Environ. Epidemiol. 5(1), 21–34.

Chapter 44

Modeling and Predicting Pesticide Exposures Daniel Vallero1, Sastry Isukapalli2, Valerie Zartarian1, Thomas McCurdy1, Tom McKone3, Panos Georgopoulos2 and Curt Dary1 1

U.S. Environmental Protection Agency, National Exposure Research Laboratory, Research Triangle Park, North Carolina Environmental and Occupational Health Sciences Institute, a joint institute of UMDNJ-RW Johnson Medical School and Rutgers University, Piscataway, New Jersey 3 Lawrence Berkeley National Laboratories, Berkeley, California 2

44.1 INTRODUCTION Models provide a means for representing a real system in an understandable way. They take many forms, beginning with conceptual models that explain the way a system works, such as delineation of all the factors and parameters of how a pesticide particle moves in the air after a spraying event. Conceptual models help to identify the major influences on where a chemical is likely to be found in the environment, and as such, need to be developed to help target sources of data needed to assess an environmental problem. In general, developing a model requires two main steps. First, a model of the domain and the processes being studied must be defined. Then, at the model boundaries, a model of the boundary conditions is especially needed to represent the influencing environment surrounding the study domain. Research scientists often develop physical or dynamic models to estimate the location where a chemical would be expected to move under controlled conditions, only on a much smaller scale. For example, the U.S. Environmental Protection Agency (EPA) uses chambers or even full-size test homes to model the fate of pesticide after application. Like all models, the dynamic model’s accuracy is dictated by the degree to which the actual conditions can be simulated and the quality of the information that is used. Numerical models apply mathematical expressions to approximate a system. There are three thermodynamic systems: 1. Isolated systems, in which no matter or energy crosses the boundaries of the system (i.e., no work can be done on the system). Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

2. Closed systems, in which energy can exchange with surroundings, but no matter crosses the boundary. 3. Open systems, in which both matter and energy freely exchange across system boundaries. Isolated systems are usually encountered only in highly controlled reactors, so are important in modeling pesticide formulation and manufacturing, but are not directly pertinent to pesticide exposure modeling. In fact, most microenvironmental systems are open, but with simplifying assumptions, some subsystems can be treated as closed. Pesticide transport and fate models can be statistical (stochastic) and/or deterministic. Statistical models include the pollutant dispersion models, such as the Lagrangian models, which follow the movement of a control volume starting from the source to the receptor locations. These often assume idealized Gaussian distributions of pesticides from a point of release; i.e., the pollutant concentrations are normally distributed in both the vertical and horizontal directions from the source. The Lagrangian approach is common for atmospheric releases, and recent models based on this approach have incorporated additional descriptions of complex turbulence. Stochastic models are statistical models that assume that the events affecting the behavior of a chemical in the environment are random, so such models are based on probabilities. These are being commonly adopted in the modeling of human exposures (see Section 44.4.1). Figure 44.1 presents a schematic of different types of modeling approaches for studying human exposures to pesticides. Deterministic models are used when the physical, chemical, and other processes are sufficiently understood so as to be incorporated to reflect the movement and fate of chemicals. Often, they are difficult to develop because 995

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Models based on physical and chemical principles

Models based on statistical relationships

Hybrid models

Pollutant concentrations in microenvironments (modeled or measured) Indoor

Time-activity pattern information

Spatial information (including Geographic Information Systems data)

Ambient

Modeled exposure FIGURE 44.1 Schematic representation of steps involved in human exposure modeling. Source: Vallero, D. (2007), adapted from U.S. National Research Council, Chapter 6: Models, in Human Exposure Assessment for Airborne Pollutants: Advances and Applications. Committee on Advances in Assessing Human Exposure to Airborne Pollutants and Committee on Geosciences, Environment, and Resources. National Academy Press, Washington, DC, 1990.

each process must be represented by a set of algorithms in the model. Also, the relationship between and among the systems, such as the kinetics and mass balances, must also be represented. Thus, the modeler must “parameterize” every important event following a pesticide’s release to the environment. In order to overcome these demands, hybrid models using both statistical and deterministic approaches are often used, for example, when one part of a system tends to be more random while another has a very strong basis in physical principles. Numerous models are available to address the movement of chemicals through a single environmental medium, but increasingly, environmental scientists and engineers have begun to develop multimedia models, such as compartmental models that help to predict the behavior and changes to chemicals as they move within and among reservoirs (e.g., carpet), in the air as dust and vapors, and in exchanges with surfaces (see the mass balance discussion in Section 44.2.1).

44.1.1 Terminology for Human Exposure Models In recent years, although an extensive array of modeling tools has been developed for supporting exposure assessments, no standardized terminology exists. A consistent terminology and essential concepts in exposure modeling need to be considered before applying exposure models. ●

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First, the science of exposure modeling is a rapidly evolving field, and the development of a standard and commonly accepted terminology is an ongoing process (see, e.g., WHO, 2004; Zartarian et al., 2005). Second, it should also be mentioned that, very often, procedures that are called exposure modeling, exposure

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estimation, etc. in the scientific literature may in fact refer to only a subset of the “complete” set of steps or components required for a comprehensive exposure assessment (e.g., contaminant transport, fate, microenvironmental accumulation, uptake by receptors). In general, since an exposure/dose calculation will need data or calculated estimates (from either external models or internal modules) to quantify different processes in the exposure sequence, the issue often becomes one of priorities as well as of semantics (so there exist environmental transport and fate models with added “exposure components” as well as exposure or risk models with “environmental fate and transport components”). Third, the process of modeling human exposures to multimedia pollutants is very often identified explicitly with population-based modeling, while models describing the specific mechanisms affecting the exposure of an actual individual (at specific locations) to contaminants are usually associated with studies focusing specifically on indoor dynamics. Finally, the concept of microenvironments is critical in developing procedures for exposure modeling. In the past, microenvironments have typically been defined as individual or aggregate locations (and sometimes even as activities taking place within a location) where a homogeneous concentration of the pollutant is encountered. Thus a microenvironment has often been identified with an “ideal” (i.e., perfectly mixed) compartment of classical compartmental modeling. More recent and general definitions view the microenvironment as a “control volume,” either indoors or outdoors, that can be fully characterized by a set of either mechanistic or phenomenological governing equations, when appropriate parameters are available, given necessary initial and boundary conditions.

Chapter | 44

Modeling and Predicting Pesticide Exposures

MENTOR

Source/Stressor Formation Chemical Physical Microbial Magnitude Duration Timing

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DORIAN

Modeling Environment for TOtal Risk studies

Disease

Dose-Response Information Analysis system

Transport/Transformation Dispersion Kinetics Thermodynamics Distributions Meteorology

Emission Inventories

Altered Structure/Function

Environmental/ Microenvironmental Characterization Air Water Diet Soil & Dust

Environmental Databases

Early Biological Effect: Biomarkers

Exposure Pathway Route Duration Frequecy Magnitude

Demographic & Activity Databases

Molecular Biochemical Cellular Organ Organism

Toxicokinetics: Biomarkers

Dose Absorbed Target Internal Biologically Effective

Urine Hair Blood Nails etc.

Cancer Asthma Infertility etc.

Edema Arrhythmia Enzymuria Necrosis etc.

Bionomic Databases

Biomarker (exposure, effect, susceptibility) Databases

Physiology Databases

Individual(s) Population Statistical Profile Community of Concern Reference Population Susceptible Individual Susceptible Subpopulations Population Distributions

FIGURE 44.2 Different processes that need to be considered in source-to-dose-effect modeling of human exposures and associated health effects and risks (from Georgopoulos, 2008).

The boundary conditions typically would reflect interactions with ambient air and with other microenvironments. The parameterizations of the governing equations generally include the information on attributes of “sources” and “sinks” within each microenvironment. This type of general definition allows for the concentration within a microenvironment to be nonhomogeneous (nonuniform), provided its spatial profile and mixing properties can be fully predicted or characterized. By adopting this definition, the number of microenvironments used in a study is kept manageable, while existing local variabilities in concentrations can still be taken into account. Microenvironments typically used to determine exposures include indoor residential microenvironments, other indoor locations (typically occupational microenvironments), outdoors near roadways, other outdoor locations, and in vehicles. Indoor residential microenvironments (kitchen, bedroom, living room, etc. or aggregate home microenvironment) are typically separated from other indoor locations because of the time spent there and potential differences between the residential environment and the work/public environment.

44.1.2 Pesticide Exposure Models A completed exposure pathway (CEP) occurs when five elements are present: a source of contamination, an environmental media and transport mechanism, a route of exposure, a point of exposure, and a receptor population (see, e.g., Georgopoulos and Lioy, 1994, 2006; Williams and

Paustenbach, 2002). The sequence of events leading from emission of a toxicant to health effects (or a biological end point in general) represents a continuum of both strongly and weakly coupled processes; exposure assessment can and should be a forward and/or backward analysis along this chain. The sequence of source-to-dose effects and major components in this sequence are shown in Figure 44.2. Human exposure to pesticides occurs when individuals contact the chemicals in various media (e.g., air, water, soil, food, dust, surfaces) through the course of their daily activities. Human exposure models, defined as a conceptual or mathematical representation of the exposure process (Zartarian et al., 2005), can be a powerful tool offering a number of advantages for human exposure assessment. For example, they allow estimation of exposure where relevant measurement data may not be available. By allowing “what if” analyses for hypothetical scenarios, they can inform risk management decisions. Through sensitivity and uncertainty analyses, models can help exposure assessors understand important factors and data needs to inform new measurements collection. Exposure models can therefore address the following questions of interest to exposure assessors: What is the population distribution of exposure, including high-end exposures to pollutants of health concern? What are exposures for susceptible subpopulations? Will the exposure cause a health effect(s) of concern? What is the pattern (intensity, duration, frequency, route, timing) of exposure? What are the key sources, routes, pathways, factors, and data needs? What are the greatest uncertainties (e.g., model uncertainty, scenario uncertainty, parameter uncertainty)? How can exposure

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FIGURE 44.3 Examples of pathways and media that need to be considered in models of human exposures to pesticides.

be effectively reduced? Was an implemented exposure reduction action effective? For these reasons, exposure models can play an essential role in the overall risk characterization and risk management process (Furtaw, 2001). Multimedia human exposure models for conducting pesticide exposure and risk assessments involve physically based algorithms that simulate transfer of pesticide from media to a person’s defined exposure surfaces (e.g., mouth, nose, skin) through various exposure routes and pathways (see Figure 44.3). Inhalation exposure equations combine pesticide air concentrations (measured or modeled) with time individuals spend in different locations. Dietary ingestion exposure is modeled by combining dietary consumption information with pesticide residues in consumed foods and beverages. Dermal (skin) exposure is estimated using surface residues and surface-to-skin transfer factors (e.g., skin surface contacted, transfer efficiency or transfer coefficient). Hand-to-mouth and object-to-mouth ingestion is modeled using hand or object residue information and exposure factors such as mouthing frequency and mouthing removal efficiency of pesticides from hands or objects. Pesticide exposure modeling assessments for regulatory purposes often involve a tiered approach. Lower tiers can include screening-level analyses with conservative default assumptions or case-specific inputs. If the screening analyses indicate the pesticide poses no problem as determined by the assessor, additional analyses may not be conducted; otherwise, higher tier analyses using probabilistic human exposure models may be conducted. This multitier process depends on various factors including available data and the type of exposure assessment required for the health end point of concern (e.g., acute, chronic). Both screening-level and higher tier models can take a microenvironmental modeling approach that combines human activity and location information and environmental concentration data, either measured or simulated. A microenvironment is defined as “surroundings that can be treated as homogeneous or well characterized in the concentration of an agent” (Zartarian et al., 2005).

44.2 TYPES OF HUMAN EXPOSURE MODELS Models of human exposure to pollutants can be classified and differentiated based upon a variety of attributes: ● ●

●

●

●

Prognostic vs. diagonostic models Population-based exposure models (PBEM) vs. specific individual-based exposure models (IBEM) Deterministic vs. probabilistic (or statistical) exposure models Observation-driven vs. mechanistic environmental model driven (e.g., with respect to providing estimates of spatially and temporally varying air pollutant concentration fields, providing estimates of contaminant levels in drinking water) Potential vs. actual exposure models (e.g., inhalation exposures models of outdoor exposure [typically maximum or potential exposures] vs. models that include locally modified microenvironmental exposures, both outdoor and indoor [actual exposures])

Estimation of exposures can be performed through predictive (or prognostic) exposure modeling studies or through statistical (diagnostic) analyses of data from field exposure studies. Prognostic assessment of exposure and dose is based on estimates of contaminant emissions, modeling of environmental transport and fate, modeling of population time/location and activity patterns, modeling of microenvironmental environmental quality, modeling of intake/uptake of contaminants, and modeling of absorption, distribution, metabolism, and elimination of the contaminant from the bodies. Diagnostic assessment of exposure and dose would be based on biomarker information, through either statistical analysis of the biomarker data or a more detailed analysis using Bayesian approaches for model-data fusion (see, e.g., Georgopoulos et al., 2009). Screening-level models for pesticide exposure assessments are typically deterministic (e.g., use point estimate

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values for all inputs) and use average estimates of time individuals spend in microenvironments (e.g., daily averages) rather than sequential time-location-activity pattern information from surveys such as EPA’s Consolidated Human Activity Database (CHAD) (McCurdy et al., 2000). These models typically focus on one exposure route/pathway at a time (e.g., inhalation, dermal contact, hand-to-mouth ingestion, object-to-mouth ingestion, food ingestion, drinking water ingestion). An example algorithm for screening-level purposes is the following: average daily hand-to-mouth exposure (mg/day)  dislodgeable residue (mg/cm2)  skin surface area mouthed (cm2/event)  hand-to-mouth contact frequency (#/h)  saliva removal efficiency (%)  exposure time (h). Screening-level models used by EPA include PIRAT (U.S. EPA, 2007), the ORD/NERL draft protocol equations (e.g., U.S. EPA, 2001), and equations included in the Exposure Factors Handbooks (U.S. EPA, 1997).

44.2.1 Screening-Level Exposure Models Screening-level multimedia mass balance models synthesize information about partitioning, reaction, and intermedia transport properties of a chemical in a representative or generic environment to assess impacts such as health risk. In a mass balance exposure model, the environment (indoor, local, urban, regional, or global) is treated as a set of compartments that are homogeneous subsystems exchanging air, particles, water, nutrients, and chemical contaminants with other adjacent compartments. Multimedia mass balance models must confront three core postulates: 1. Conservation of mass: Chemicals put into the environment will accumulate, particularly if removal processes are slow, but the capacity of any environmental systems to assimilate chemicals is finite. 2. Chemical equilibrium: The second law of thermodynamics tells us that chemicals will be distributed in environmental systems in a way that minimizes the free energy of that system. The change in free energy associated with movement of the solute from one compartment to another is directly proportional to the difference in chemical potential between the compartments. 3. Integrated systems models are best suited to capture conservation of mass and chemical equilibrium: the environment can be usefully described as a set of linked compartments or boxes. The underlying assumption is that the overall fate of a chemical of interest is more strongly controlled by partitioning between the various phases available to the chemical than by spatial differences in properties within individual compartments of the system. This postulate is consistent with a philosophical approach to model development that views a useful model as one that captures the characteristics of a system that are assumed to be important and omits those that are assumed to be extraneous.

999

With these three postulates as a foundation, multimedia mass balance models describing the partitioning and ultimate fate of chemicals in the environment can be assembled.

44.2.1.1 Fugacity Models and Fugacity Capacities Fugacity is a metric for quantifying chemical potential at low concentrations. Fugacity, f, can be viewed as the “escaping tendency” of a chemical in a phase, has dimensions of pressure, and is related to concentration, C, by a proportionality constant, fugacity capacity, Z: C  Zf

(1)

In the SI system, Z has units of mol/m3 Pa. From the ideal gas law, it can be shown that Z for the vapor phase is 1/RT where R is the gas constant and T is absolute temperature in kelvins. Fugacity relates to equilibrium, but can also be used to explore systems out of equilibrium. When a chemical reaches equilibrium distribution between two available phases, the fugacities of the chemical in the phases are equal. Equilibrium partitioning between two phases can also be described by a dimensionless partition coefficient K12, which can be measured under laboratory conditions as the ratio of concentrations C1 and C2. Applying the relationship between concentration and fugacity, and recognizing that f1  f2 at equilibrium, K12 

C1 f Z1 Z   1 C2 f Z2 Z2

(2)

Z relationships can be determined experimentally for many phases by measuring partition ratios between the phase of interest and a phase with known Z. Table 44.1 illustrates how one obtains Z values in a mass balance model that includes water, soil, and sediment phases by using the Z value for air as a starting point and appropriate partition coefficients measured in the laboratory. Several assumptions are required to arrive at these expressions, including (1) that chemicals in the vapor phase obey the ideal gas law, (2) that chemicals in the aqueous phase form ideal dilute solutions, and (3) that octanol can be used as a surrogate to describe chemical partitioning to lipids and the organic carbon component of soil and sediments (Mackay, 2001; MacLeod and McKone, 2004).

44.2.1.2 Diffusive Transport in Mass Balance Models In a fugacity model, the net diffusive flux, in mol/m2 d, of chemicals across the surface area separating two compartments (for example, air and surface soil) is flux  Y12 ( f1  f2 )

(3)

where Y12 is the fugacity mass-transfer coefficient across the boundary between compartments 1 and 2 with units mol/(m2 Pa d) and f1 and f2 are the fugacities of compartments 1 and 2.

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For example, the flux in mol/d/m2 of a pesticide from air to surface soil through particle deposition is the product of the particle deposition velocity, vd in m/d, the fugacity capacity of the pesticide in air particles, Zap, and the total fugacity of the bulk air compartment, fa:

TABLE 44.1 Fugacity Capacities (Z Values) for Environmental Phases Phase

Definition of Z (mol/m3 Pa)

Air

ZA  1/(RT )

R  8.314 Pa m3/mol K T  temperature (K)

Water

ZW  1/H  CS/PS

H  Henry’s Law constant (Pa m3/mol) Cs  aqueous solubility (mol/m3) Ps  vapor pressure (Pa)

Soil solids

ZS  ZW  yS  k  yS  fraction of organic KOW  ρS/1000 carbon in soil k  Karickhoff constant  0.41 l/kg KOW  octanol–water partition coefficient ρs  density of soil (kg/l) 1000 converts l to m3

Sediment solids

Zd  ZW  yd  k  KOW  ρd/1000

flux (air to ground-surface soil)  vd  Z ap  fa

Examples of intercompartmental advection processes are rainfall, deposition of atmospheric aerosol particles, resuspension of particles from soil, water-borne erosion of soil, runoff of precipitation, infiltration of water through soil, deposition of sediment particles in surface water, resuspension of sediment particles from the sediment layer, and surface water flows. Once equations describing all intercompartmental transfers of contaminants have been derived, mass balance equations equating input and removal rates can be written for each compartment (i) of the environmental system. Ei  ∑ flow j → i 

yd  fraction of organic carbon in sediment ρd  density of sediment (kg/l)

Equation (3) is analogous to the flow of heat in a system where ( f1  f2) plays the role of a temperature difference, Y12 is heat conductance, and the mass flux is the equivalent of heat flux in J/(m2 d). The fugacity mass-transfer coefficient depends on the mass-transfer coefficient on either side of the interface and the fugacity capacities of the two media that form the interface. 1

⎛ 1 1 ⎞⎟ ⎟⎟  Y12  ⎜⎜⎜ ⎜⎝ Z1U1 Z 2U 2 ⎟⎠

(4)

where U1 and U2 are the mass-transfer coefficients (m/d) in the boundary layers in compartments 1 and 2 and Z1 and Z2 are the fugacity capacities of compartments 1 and 2.

44.2.1.3 Advective Transport in Mass Balance Models The intercompartment transfer of chemicals by advection is also modeled as a flux at the surface between two compartments, such as air and soil. To be consistent with the area normalized description of diffusion given previously, this flux (mol/m2/d) is modeled as the product of the velocity of the moving phase (m/d) and the contaminant concentration in that phase (mol/m3). Advection flux  velocity  Z ik fi

(6)

(5)

where Zik represents the fugacity capacity of the moving phase and fi the fugacity of the pesticide of interest in compartment i.

∑ flowi → j  ∑ flowi → sin k

(7)

On the left-hand side of Eq. (7) are chemical inputs to compartment i by direct emission, Ei, and the total rate of intercompartmental transfer to compartment i, Σflowj→i. Removals from the compartment occur by intercompartmental transfer (Σflowi→j) and by advection out of the system or chemical transformation (Σflowi→sink). For an environment consisting of n compartments, one can write n equations of this type and solve them algebraically to obtain the fugacity of pesticide in each compartment.

44.2.1.4 Complexity and Level of Detail Mackay (2001) has noted that multimedia mass balance models have four levels of complexity based on the whether a system is open or closed, in equilibrium or not, and steady state or dynamic. These are levels are as follows: Level I: closed system, equilibrium, and steady state; Level II: open system, equilibrium, and steady state; Level III: open system, nonequilibrium, and steady state Level IV: open system, nonequilibrium, and dynamic In addition to complexity with respect to system boundaries, chemical equilibrium, and time, mass balance models can be organized at different levels of spatial detail. There is the issue of both what geographic scale is being represented – an indoor environment, a neighborhood, an urban area, a region, or even the earth system – and the spatial detail used to represent that geographic scale. For example, a mass balance model tracking pesticide transport for an indoor environment might be organized into a systems model with a single air volume, floor surface, wall surfaces, and furniture to represent the household system or it could have this type of structure applied to subcomponents

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FIGURE 44.4 Generalized schematic of a regional mass balance model showing gains and losses and the transfer of mass driven by differences in fugacity among the moving and nonmoving phases in an agricultural region such as the Salinas Valley of California.

FIGURE 44.5 Generalized schematic of a steadystate indoor mass balance model showing gains and losses and the transfer of mass driven by differences in fugacity among the moving and nonmoving phases.

representing different rooms or groups of rooms. Similarly, a mass balance model tracking the fate of pollutants used in an agricultural region such as the Salinas Valley of California might be organized with single air, soil, surface water, and sediment compartments representative of the whole region or be divided into two or more subregions with air, soil, surface water, and sediment compartments linked among adjacent subregions. Figure 44.4 illustrates a regional mass balance model for tracking pesticide fate in an agricultural region. Figure 44.5 illustrates a mass balance model that organizes the indoor environment into a single air volume and surfaces.

44.2.1.5 Mass Balance Models of the Ambient Environment Multimedia mass balance models are used widely for screening-level chemical assessments, for cumulative assessments of pesticide exposures, for assessing the regional and global fate of persistent organic chemicals, and for life cycle

impact assessment. Some multimedia fate and exposure models are based on environmental parameters that are not representative of any specific geographical area. These generic models are used as a “laboratory” for evaluating the likely behavior of pollutants and how this relates to basic chemical properties. Generic multipathway human exposure models coupled with multimedia fate models have been developed in the United States, Canada, and Europe. Recently, generic models of contaminant fate have been adapted to conduct rapid screening-level assessments of large numbers of chemicals for persistence (P) and potential for long-range transport (LRT). Regional multimedia mass balance models have the same framework as generic models but include geographical databases representing a specific political or ecological region. Multiregion models include more spatial resolution by linking several regional mass balance models and have been applied on local, continental and global scales. An advantage of regional models over evaluative models is that results can be directly compared with reported concentrations of contaminants in a specific area.

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44.2.1.6 Mass Balance Models of the Indoor Environment As originally developed, indoor mass balance models focused on air volumes and tended to ignore the retention and re-emission of chemicals by surfaces. In some cases, particularly for pesticides, empirical models were fit to observations of the time history of air concentrations following chemical application. But these approaches fail to address the role of indoor surface retention capacity in controlling the long-term behavior of chemicals brought to the indoor environment. Bennett and Furtaw (2004) proposed a fugacitybased indoor pesticide model that is now widely used. This is a dynamic mass balance model with several compartments simultaneously exchanging mass. The room or rooms where the pesticide is applied are the pesticide-treated zones and adjacent rooms comprise one or more untreated zones. Each room or zone includes an air compartment and three surface compartments – carpet, vinyl floor, and walls/ceiling. They assume a specified thickness for each surface material, and further assume that the compound is evenly distributed through the surface material. They define the mass balance equations, the fugacity capacity of each compartment, and the transfer rates among the compartments.

44.2.2 Comprehensive Human Exposure Models In recent years, an extensive number of comprehensive modeling tools were developed by various organizations to support quantitative exposure analyses of pesticides. As part of the development effort for the TRIM (total risk integrated methodology) modeling system, the U.S. EPA conducted an extensive review of models, available at the time, that were considered relevant to exposure assessments. These models typically focus on a particular process or on certain components of the source-to-dose sequence. Furtaw (Furtaw, 2001) summarized developments regarding Models-3/Community Multiscale Air Quality (CMAQ) model, Stochastic Human Exposure and Dose Simulation (SHEDS) (Burke et al., 2001; Zartarian et al., 2000, 2006), Exposure Related Dose Estimating Model (ERDEM) (Blancato et al., 2004; U.S. EPA, 2009a) and FRAMES-3MRA (Babendreier and Castleton, 2005; U.S. EPA, 2008b). “Person-oriented” modeling and multimedia models for aggregate and cumulative exposures (including Lifeline, CARES, and Calendex) have been reviewed by Price et al. (2003). Parallel efforts toward system integration for incorporating, in a consistent framework, both environmental and biological models are being pursued and implemented, including the MENTOR (Modeling Environment for Total Risk) studies system. The focus of this effort is a library of software modules or a “computational toolbox” intended to facilitate consistent multiscale source-to-dose modeling of

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exposures to contaminants for individuals and populations. This is achieved by linking predictive models of fate/transport and exposure/dose, coupled with up-to-date national, regional, and local databases of environmental, microenvironmental, biological, physiological, demographic, etc. parameters. In addition to traditional components for modeling human exposure, MENTOR includes sets of novel computational tools for systematic model sensitivity/uncertainty analysis, model reduction, data mining via pattern recognition, etc., which are valuable for exposure reconstruction. A multimedia implementation of MENTOR, called MENTOR-4M (Modeling ENvironment for TOtal Risk studies for Multiple co-occurring Contaminants and Multimedia, Multipathway, Multiroute Exposures (4M)), is especially relevant to modeling pesticide exposures (CREM; U.S. EPA, 2008a). The multimedia MENTOR implementation currently considers five exposure pathways: inhalation, drinking water consumption, food intake, nondietary ingestion, and dermal absorption. These pathways have been implemented in both individual-based exposure modeling (IBEM) and population-based exposure modeling (PBEM) approaches. Both these approaches employ a person-oriented modeling (POM) formulation, i.e., they are driven by the attributes and activities of the exposed real and/or virtual individual(s). While IBEM implementations utilize the information relevant to actual individuals (and produce exposure and dose estimates specific for each one of them), the PBEM implementations focus on the statistical characterization of the exposures and doses of selected populations (at the level of census tract, county, state, etc.). Thus, the questions posed by any particular environmental health problem can be tailored to small sets of individuals potentially at risk or to larger populations or subpopulations of interest.

44.3 SOURCE-TO-DOSE EXPOSURE MODELING FOR PESTICIDES Categories of inputs for pesticide exposure models include simulated population information (e.g., demographic factors, time-location-activity pattern information, dietary consumption information), pesticide-specific data (e.g., measured or modeled residue and concentration data for various media, decay rates, usage and application information), and other exposure-related factors (e.g., mediato-person pesticide transfer factors). Outputs can be in the form of time-averaged exposure for short-term, intermediate, or longer-term (e.g., lifetime) exposure durations. Some of the microenvironmental models mentioned previously are calendar based and can preserve sequential exposure time profiles using a minute, hour, or day as the time step for calculating exposures. Various metrics can be extracted from these time profiles for Monte Carlo simulations, or the profiles that preserve sequence and patterns of exposure (by an individual pathway or aggregated across

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FIGURE 44.6 Integrative modeling of source-to-dose exposures for individuals and populations. Processing occurring at various scales, across different environmental media, along with individual-specific factors (activity, physiological attributes, etc.) impacts the final exposures experienced by specific individuals and populations. Left panel from 3MRA User Guide 2002; right panel from Georgopoulos et al., ES&T, 1997, 31(1).

pathways) for a simulated individual can be provided as inputs to dose estimation models (e.g., toxicokinetic models). Figure 44.6 presents a schematic description of the processes and scales involved in the source-to-dose modeling for assessing human exposures to pesticides.

44.3.1 Overview of the Steps in Exposure Modeling For assessing population exposures to multimedia environmental chemicals, the person-oriented PBEM framework (summarized schematically in Figure 44.7) provides the following seven steps that consider inhalation, drinking water consumption, food intake, and nondietary ingestion exposure routes. These steps are general in nature and can together be considered as a template for conducting prognostic exposure modeling for multimedia chemicals: 1. Estimation of the multimedia background levels of pesticides (in air, water, soil, food, etc.), for the area/locations where the population of interest resides, through data from outcomes of environmental models and/or from field studies. A wide range of environmental models exists for studying fate and transport of chemicals in different media (such as Models-3/CMAQ CB4 and CBTOX versions (Georgopoulos et al., 2005; Luecken et al., 2006; U.S. EPA, 1999), CAMx (ENVIRON, 2004), AERMOD (U.S. EPA, 1998), CALPUFF (Scire et al., 2000), MM5 (NCAR, 2004), RAMS/HYPACT (Walko and Tremback, 2001; Walko et al., 1999) for air pollutants; EPANET2 (Rossman, 2000) for water-borne contaminants in municipal networks;

MODFLOW (Guiguer and Franz, 1996), FACT (Hamm and Aleman, 2000) for groundwater contaminant transport; CATS (Fascineli et al., 2002; Traas et al., 1996) for foodweb simulations; FRAMES/3-MRA (U.S. EPA, 2003) for multimedia fate and transport simulations, etc.). For exposure assessment purposes, ambient pollutant concentration information available at a local level (such as census tract or neighborhood) may be needed as an input to microenvironmental models for the estimation of population or individual exposures. However, typical field monitoring networks as well as regional environmental quality models provide spatial characterizations of concentration fields that often are too coarse for exposure characterization, so there is a need to further characterize local variability. This can be accomplished through spatio-temporal random field (STRF) (Christakos and Vyas, 1998), as well as Bayesian maximum entropy (BME) (Christakos and Serre, 2000). 2. Estimation of multimedia levels (indoor air, drinking water, soil/dust, food, etc. concentrations) and temporal profiles of environmental contaminants in various microenvironments such as residences, offices, restaurants, and vehicles. These can be calculated through microenvironmental steady-state or dynamic mass balance model simulations (described earlier), supplemented by information from empirical databases. 3. Selection of a fixed-size sample population of “virtual individuals” in a way that statistically reproduces essential demographics (age, gender, race, occupation, education) of the population unit used in the assessment (e.g., a sample of 500 virtual individuals is typically used to represent the demographics of a given census tract,

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i.a. Databases: AIRS, NET, NATA, CEP, WQN, NAWQA, STORET, EMAP, NGA i.b. Models: CMAQ, REMSAD, AERMOD, ASPEN, GMS, FACT, MODFLOW, WMS, CATS 1. Estimate multimedia background levels of pesticides (in air, water and soil) through either: a. multivariate spatiotemporal analysis of measurements b. regional-scale environmental model predictions Calculate Potential Outdoor Exposures

3. Characterize attributes of populations (geographic density, age , gender, race income, etc.) a. select fixed-size sample population that statistically reproduces essential demographics or b. divide population of interest into exhaustive set of cohorts

US Census, US Housing Survey, Local Data

ii.a. Databases: SDWIS/FED, TDS, PDP, CSFII NHEXAS, NHANES ii.b. Models: APEX, HAPEM, SHEDS-Multimedia, EPANET, DEPM, DEEM

2. Estimate local pesticide levels in an administrative unit (such as a census tract) or a conveniently defined grid through either: a . field study measurements b. subgrid “corrections” of regional model estimates c. application of a local scale environmental model

4. Develop activity event (or exposure event) sequences for each member of the sample population or for each cohort for the exposure period through either: a. existing databases from composites of past studies (for baseline assessment) b. study-specific information (special registries)

CHAD, NHAPS

5. Estimate multimedia levels and temporal profiles of pesticides in various microenvironments (outdoor, indoors, etc.) through either: a. field study measurements b. microenvironmental massbalance model (air), drinking water distribution model (water), dietary exposure model (food)

6. Calculate appropriate inhalation rates, as well as drinking water and food consumpation rates for the members of the sample population combining the physiological attributes of the study subjects and the activities pursued during the individual exposure events

Calculate Exposures/ Intakes

7. Biologically Based Target Tissue Dose Modeling

ICRP and Other Physiological & METS Databases, CSFII, NHANES

FIGURE 44.7 An overview of the steps involved in individual or population-based modeling of human exposures to pesticides (adapted from Georgopoulos and Lioy, 2006).

and a sample of 10,000 virtual individuals is typically used to represent the demographics of a county). The population attributes, such as the distributions of age, gender, employment, and housing, can be developed from available census data (e.g., from the U.S. Census Bureau (USCB, 2009) or from study-specific definitions). Sometimes, relevant databases are available as components of other modeling systems, as in the case of the Air Pollution Exposure Model (APEX) (Glen, 2002), which provides databases for housing as well as for commuting profiles. 4. Development of activity event (or exposure event) sequences for each member of the sampled population or for each cohort for the exposure period through one of the following: a. Retrieval of matching time-activity diary records from existing databases from composites of past studies (e.g., from U.S. EPA’s CHAD (McCurdy et al., 2000; Stallings et al., 2002)) for each virtual individual of the sample population, based on each individual’s demographic characteristics. b. Hypothetical scenario-based or “simulated” activity patterns based on study-specific needs. CHAD (McCurdy et al., 2000; Stallings et al., 2002) contains over 22,000 person-days (diary records) of activity patterns developed from pre-existing human activity studies. Each diary record provides a basis for simulating the movement of the virtual individual through geographic locations and microenvironments during the simulation period. Each event is defined by

geographic location, start time, duration, microenvironment visited, and an activity performed. The attributes of CHAD records include age, gender, employment status, and smoking status of each individual, which can be used for matching the demographic characteristics of each sampled virtual individual. For additional analysis of hypothetical exposure scenarios, activity profiles can be synthesized either independently or through scenario-specific modifications to existing CHAD diaries. Activity-based exposure modeling is described in more detail in the following subsection. 5. Estimation of multimedia levels (indoor air, drinking water, and food concentrations) and temporal profiles of multimedia chemicals in various microenvironments (residences, offices, restaurants, vehicles, etc.): a. Residential indoor air concentrations can be calculated using microenvironmental mass balance modeling with inputs from step 2. Nonresidential microenvironments (office, school, restaurant, etc.) can be developed either through mass balance modeling or through linear regression equations developed from analysis of concurrent indoor and outdoor measurement data available for the multimedia chemicals in these microenvironments. The SHEDS model (Burke et al., 2001) provides distributions of air exchange rates for different types of residential microenvironments, while other models and databases provide distributions for air exchange rates for general nonresidential microenvironments (Turk et al., 1989) and vehicle microenvironments (Hayes, 1991).

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b. Drinking water concentrations can be obtained from regulatory monitoring databases (such as SDWIS/ FED; U.S. EPA, 2008c) or field study measurements (such as National Human Exposure Assessment Survey (NHEXAS); Pellizzari and Clayton, 2006; Thomas et al., 1999). If such data are not available, the drinking water distributions can be modeled using drinking water models such as the EPANET2 model (Rossman, 2000) with treatment plant data to obtain drinking water concentrations (see, e.g., Maslia et al., 2000). c. Food concentrations can be obtained from survey studies such as the Total Diet Study (TDS) (Tao and Bolger, 1999) and NHEXAS (Pellizzari and Clayton, 2006; U.S. EPA, 2006b).

(Peterson et al., 2001; Tao and Bolger, 1999), covering 1991–1999, provides information on average total concentrations of several multimedia chemicals in 267 types of raw agricultural commodities, which are composites of food items. Recipe files for specific types of food intakes can be developed and linked to the CSFII and TDS databases to generate the estimates of dietary intakes of different chemicals. d. The magnitude of nondietary intake of individual chemicals from incidental soil/dust ingestion can be estimated using age-specific empirical intake rate distributions fitted to available tracer element mass balance study results. The distributions of estimated soil and dust ingestion rates can be obtained from surveys such as those from Buck et al. (2001).

6. Calculation of inhalation and ingestion intake (drinking water, dietary, and nondietary) rates for the members of the sample population, reflecting/combining the physiological attributes of the study subjects and the activities pursued during the individual exposure events.

These estimates can then be apportioned by using times spent outdoors vs. indoors into corresponding soil and dust ingestion exposures.

a. The drinking water intake rates can be estimated by extracting appropriate survey records (from, e.g., the Continuing Survey of Food Intake by Individuals database (CSFII); Tippett et al., 1999) matching the virtual individual’s demographic characteristics. These rates include (1) consumption of tap water directly for drinking, (2) amount of tap water used in food and home-prepared cold beverages (e.g., lemonade mixes), and (3) hot beverages (e.g., coffee, tea). It should be noted that currently available studies on drinking water intake are based on short-term survey data and may have certain limitations, especially with respect to upper percentile values. The CHAD diary records also provide information on energy expenditure, which can be used to estimate water intake rates. b. The inhalation rate can be calculated based on the person’s age, gender, and the metabolic equivalent of tasks (METS) value associated with the activity pursued (see e.g., Georgopoulos et al., 2005). The energy expenditure information from CHAD diary records can be used directly to estimate inhalation rates. Alternatively, probability distributions or tables describing age-specific inhalation rates of humans can also be used (see, e.g., Allan and Richardson, 1998; Brochu et al., 2006a,b). c. Dietary intake of multimedia chemicals for each virtual individual is estimated utilizing the following information: food consumption rates, composition of food item (recipe file), and arsenic residue data in food. The U.S. Department of Agriculture’s (USDA’s) CSFII database provides information on food consumption rates for the general U.S. population, covering 1994–1996 and 1998. The U.S. Food and Drug Administration’s (USFDA’s) TDS database

7. Combination of each virtual individual’s inhalation and ingestion intake rates with the corresponding microenvironmental concentrations of chemicals of interest, for each activity event and location, to assess exposures and estimation of target tissue doses of arsenic and its metabolites through physiologically based toxicokinetic modeling (PBTK modeling). The use of PBTK modeling should be contrasted with commonly used assumptions that use linear relationships for relating the internal dose to cumulative exposure through a given pathway and exposure route. Simple pathway exposure factors are often used as a result of this assumption to relate ambient concentration to dose. These assumptions can be violated due to nonlinear processes of metabolic elimination, saturation effects, etc. (Smith, 1992). Furthermore, the internal and target tissue doses due to exposure to the same agent via different routes (e.g., ingestion and inhalation) are not necessarily additive, and care must be taken in defining total dose due to multiroute exposures. Thus, the most scientifically sound approach for relating exposure to internal dose would be provided through PBTK models that can account for the distribution and metabolism of the chemical in the various organs. Use of PBTK models in exposure assessment should consider the applicability and limitations of linearity assumptions for exposure-dose, the contaminants and time scales of exposure that require explicit treatment of the effects of temporal variation of exposure on total dose, and the evaluation of the applicability of existing modeling tools for relating – either statically or dynamically – exposure to dose. PBTK models can provide the basis for evaluating assumptions related to exposure-dose linearity, physiological damping of exposure, etc. A wide variety of computational tools exists for pharmacokinetic (PK) and PBPK modeling (Bonate, 2005;

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Gibaldi and Perrier, 1982; Isukapalli et al., 2007; Jacquez, 1972, 1985; Reddy et al., 2005; U.S. EPA, 2005). The MENTOR-3P and ERDEM systems allow development of PBPK models for specific chemicals or groups of chemicals through the formulation of “libraries” that can be refined and shared among scientists. These libraries can incorporate literature data in consistent frameworks. Applications of PBTK modeling in the context of occupational exposures are discussed in Dary et al. (1996) and Knaak et al. (2002).

EPi  ( E1 , E2 , E3 … En ) for time period T

44.3.2 Human Activity-Based Exposure Modeling According to the EPA, there are three techniques to estimate pollutant exposures quantitatively. Sometimes the approaches to assessing exposure are described in terms of direct measures and indirect measures of exposure (e.g., NRC, 1994). Measurements that actually involve sampling on or within a person, for example, use of personal monitors and biomarkers, are termed direct measures of exposure. Use of models, microenvironmental measurements, and questionnaires, where measurements do not actually involve personal measurements, are termed indirect measures of exposure. The direct/indirect nomenclature focuses on the type of measurements being made; the scenario evaluation/point-of-contact/reconstruction nomenclature focuses on how the data are used to develop the dose estimate. The three-term nomenclature is used in these guidelines to highlight the point that three independent estimates of dose can be developed. There is seldom a sufficient amount of direct information from measurements to estimate the distribution of exposures in a population, which is most useful for understanding health risks associated with exposures. Thus, an indirect method must be used. An instantaneous personal exposure can be expressed mathematically as a composite of pollutant concentration in a medium that comes in contact with a target of interest – a person – at a given instant of time (Duan, 1991). Et  C t * t

(8)

where Et is exposure for time period t (example units: μg/min), Ct is the steady-state concentration of a substance of interest (μg), and t is the time period that a target comes in contact with Ct (min). Dose is the amount of pollutant taken into the body during an exposure; it is known either as the uptake or intake dose rate (IDR) depending upon the media of exposure. We will use IDR for this section. It is the amount of material passing through the skin or entering the lung per unit time period. Dt  Et * IDR

where Dt is the dose rate for time period t (example units: moles/min). Ct in a particular location varies over the course of a day or longer time period, T. Thus, it is function of the particular period of time spent in a particular location by a person. When Ct is steady over a particular time interval, the location is called a microenvironment (μE) by definition. People go in and out of different μEs over T, and the subsequent exposure and dose “profiles” are captured thus:

(9)

(10)

where EPi is the exposure profile for individual I for time T. Similarly, the dose profile for each individual is defined to be DPi  ( D1 , D2 , D3 ,… Dn ) for time period T

(11)

If the person spends time in a μE without a concentration of interest, the exposure and dose estimates for that time period are 0. An integrated exposure and dose estimate for T for an individual is T

EiT 

∫ EPi

(12)

and T

DiT 

∫ DPi

(13)

The average exposure and dose for an individual for time period T is simply the appropriate integral divided by 1/T. This logic can be extended to a population by integrating across people, but since modelers are interested in the distribution of exposures and intake dose across a population, there is not much sense in doing so. It is the distribution of exposure/dose profiles that gives rise to population health risks. Application of this logic requires measurements (or modeled estimates) of concentrations or the sequence of concentration values. When observations are used, a concern not always addressed is the use of concentrations below the detection limits of the measuring instrument. A further complication is that of detectable but low background concentrations, which can be a critical factor in determining both the duration and the overall magnitude of integrated exposures. This issue is especially important when aggregate multimedia exposures from multiple pathways result from low concentrations in each medium, each of which may be below the detection limit, but the combined exposures can result in significant amounts of dose.

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Modeling and Predicting Pesticide Exposures

When the concentration estimates reflect an average over a time period, actual information may be lost because the time averaging process serves as a low-pass filter. This can be critical if fast and/or nonlinear phenomena are present somewhere in the exposure sequence (e.g., in environmental transformations). Many averaging/filtering and lumping/approximation methods commonly used to assess exposures through measurement and modeling or a combination of both may introduce substantial errors in this assessment. Although these procedures cannot be avoided in many analyses of practical situations, it is essential that their effects be considered, at least via an order of magnitude analysis, prior to conducting an exposure assessment. This will ensure that the exposure characterization is not driven by invalid assumptions. Furthermore, exposure assessment should focus explicitly on individuals (person-oriented modeling) and not on exposure locations (points, areas, microenvironments, etc.) per se. This represents a major challenge, as very few studies measure personal concentrations. What is routinely monitored is often ambient concentration; indeed, numerous databases are available that archive information from fixed location monitors and this information must be used to derive, via modeling, the personal exposure concentration profile. Modeling where a person is in time and space requires a database on time/activity information, also known as time budget data. Indoor and personal exposures often dominate exposures to many pesticides. Proper characterization of the time a person spends at various activities in each microenvironment is essential for comprehensive exposure modeling. In general, time use data must include attributes

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for location, specific activities and activity level. Of these, location is the most important consideration in exposure assessment. Next, activity level is the most important factor needed to estimate intake dose rate, which is needed to characterize pesticide risks. Figure 44.8 presents an overview of activity-based exposure modeling for populations. The general steps involved in generating a specific virtual individual are highlighted in Figure 44.9. These virtual individuals are generated in a manner that ensures that the statistical properties of the virtual individuals are close to those of the real individuals being studied. Human activity data are generally available from a number of sources, and are being incorporated into CHAD (McCurdy et al., 2000). Research is currently under way to improve and expand the amount of human activity data in CHAD, including the following: ●

●

● ●

●

obtaining additional personal diaries for input into CHAD, particularly those with multiple days of data; analyzing the American Time Use Survey (ATUS) data for their applicability and suitability and exposure assessments; evaluating and improving commuting methodologies; developing approaches for better longitudinal diaries by addressing intra- and interindividual relationships in the U.S. population; and developing CHAD-Explorer.

A crucial aspect of activity-based modeling is the need to simulate the person who will be undertaking the activities. As shown in Figure 44.8, the simulated person exists in space and time. From these parameters, diary results are

FIGURE 44.8 Schematic description of human activity-based exposure modeling for populations.

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1 6

Simulated Individual • Home location • Work location (if employed) • Age • Gender • Ethnicity • Employment status • Housing characteristics • Anthropometric parameters (height, weight, etc.) • Basal Metabolic Rate (BMR)

Individual Physiological Sequence Metabolic Equivalents (METS) Oxygen Consumption Rate (VO2) Total Ventilation Rate (VE) Alveolar Ventilation Rate (VA) PAI, actual daily estimate 3 Simulated Individual Activity Profile • Selected diary records days in simulation period • Sequence of events (microenvironments visited, minutes spent, and activity)

2 Activity Diary Pools • Personal attributes • Day-type (e.g., weekday) • Temperature • Physical activity index (PAI) (initial/median estimate)

5 Stochastic Calculation • Energy expended per event and ventilation rates • Both adjusted for physiological limits and EPOC

4 Physiological Parameters • METSMAX, METSRES • Ventilation relationships

FIGURE 44.9 Steps in building a realistic “virtual individual.” The objective is to ensure that the statistical properties of the population of virtual individuals are close to those of real subjects being modeled.

pooled, from which the activity profile can be developed for the simulated person. This profile is combined with physiological parameters and probability-based information about expended energy and ventilation rates to arrive at an individual physiological sequence (e.g., breathing and metabolic rates). From this sequence, a “realistic person” is used to construct the model. Developing longitudinal activity patterns from crosssectional data can be challenging. It is important not to wash out specific information. For example, a good model must maintain sufficient intraindividual and interindividual variability of the various locations. This is highly dependent on the way longitudinal activity data are aggregated from individual diaries (Xue et al., 2004). In addition, correlations in individuals according to time spent in selected locations must be maintained. Finally, individuals must be accurately classified according to lifestyle (as indicated from diary entries). In particular, sedentary individuals must be distinguished from active persons (i.e., exercisers). Further, age and gender differences may need to be noted. The differentiations can be especially powerful if combined with location (e.g., outdoor children, indoor exercisers). Modeling children’s activities is often necessary in pesticide exposure scenarios. The number of days needed to reliably estimate the amount of time children spend outdoors is an important modeling parameter. Xue et al. (2004) recommend 7 consecutive days of activity data in four seasons in order to adequately model exposures to children. Health effects are associated with the time pattern of dose rate received. Also, temporal order is important since human activities and individual physiological parameters

are correlated in time. Intra- and interindividual variability in activity along with physiological processes must be modeled and longitudinal activity relationships must be developed from cross-sectional data. Diary days are assembled for longitudinal assessment using the following steps (Glen et al., 2008): 1. ranking individual diary days (generally cross-sectional) on a user-defined metric; 2. defining numeric goals to be attained in the modeling simulation so that the observed interclass correlation coefficient (ICC) for intra-and interindividual seen in the populations is attained; and 3. defining the “lag 1” (day-to-day) correlation goal to be attained (obtained from the literature). Thus, combining activity information with physiological data and models is presently an important research focus and will help pesticide modelers to evaluate human activity data suitable for time series exposure and intake dose modeling.

44.3.3 Toxicokinetic Modeling for Assessing Pesticide Uptake and Distribution PBTK models hold the promise of linking exposure with evidence of exposure as expressed through the detection and measurement of biomarkers of exposure (Andersen et al., 2005; Blancato et al., 2004; U.S. EPA, 2006a) and have been well chronicled (e.g., Reddy et al., 2005). They describe the transport and metabolism (transformation) of chemicals

Chapter | 44

Modeling and Predicting Pesticide Exposures

within physiological systems; these are key processes affecting concentrations of chemicals in tissues, and hence the responses of physiological systems to exposure. PBTK models can be used to predict concentration-time profiles in several tissues of interest, because they are mechanistic. More specifically, PBTK models describe the body in terms of compartments that represent organs or tissues or groups of organs and tissues. The structure of a PBTK model is often derived from basic anatomical and physiological structure of the organism studied. These models are also more amenable to different types of rational, mechanistically based extrapolation, including cross-species, cross-tissue, cross-chemical, and high-dose to low-dose extrapolations. The cross-species extrapolation of PBTK models is based on the rationale that the whole body structure is similar in different mammals, such as mice, rats, dogs, and humans. In fact, a majority of PBTK models in toxicology and risk assessment have been developed by utilizing animal experiments and have been scaled for risk assessment studies in humans. PBTK models are mechanistically based and have physically meaningful parameters, most of which can be obtained from independent experiments or from the literature. A majority of remaining PBTK model parameters can be estimated from in vitro data without the need for data from expensive in vivo studies. Overall, PBTK models can provide insight into the several aspects associated with the kinetics of a pesticide within the human body, collectively termed ADMET, for absorption, distribution, metabolism, elimination, and toxicity. The mathematical formulation of a PBTK model is dependent on several factors: routes of intake of a chemical, target tissues of interest, physiological components to be explicitly modeled (kinetically important tissues and organs and the linkages between them), transport processes of the chemical (flow, diffusion, disposition, clearance, etc.), and metabolic processes involved. At a minimum, the PBTK model is expected to be representative of the biological system of interest to the point of acting as an in silico mimic such that exposure to a substance may be accurately and precisely assessed. This condition requires that the model be as complete as possible and adaptable to the exposure conditions and dose metrics that define the hazard and risk assessment. To this extent, a representative and complete PBTK model should be adaptable to any chemical or class of chemicals where ADMET parameters and structures can be “turned off and on” as needed to address the application. The model must be robust enough to extend beyond single chemical applications. The model must hold up to interactive scrutiny and not be merely printed and published. We conceptualize PBTK models as the pivot point or fulcrum balancing the weight of evidence for exposure on one side and the disposition of the exposure dose on the other. The main routes of intake of a pesticide are ingestion (dietary or nondietary), inhalation, and dermal absorption. Therefore, corresponding organs or tissues usually need to be explicitly modeled in order to describe the uptake of the

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pesticide. Organs that are usually explicitly described in a PBTK model include liver (primary site of biotransformation), lung (for volatile and semivolatile chemicals that are absorbed by inhalation and eliminated by exhalation), skin (for exposure scenarios that include dermal absorption), gastrointestinal tract (for absorption of ingested chemicals), and kidney (for renal excretion). Both arterial and venous blood streams are usually modeled separately (either explicitly as compartments or implicitly) or are linked to other compartments in a manner representative of body blood circulation. The remaining tissues and organs are usually lumped into three compartments according to their kinetic characteristics: rapidly perfused, slowly perfused, and fat. Figure 44.10 presents an example of how a whole body and intake routes are represented in a PBTK model. The membranes of interest for PBTK modeling of pesticide exposure are skin, the epithelial lining of the GI tract, and the cells and tissues of the respiratory tract and lung. These membranes define the routes of exposure as the termini of the dermal, ingestion, and inhalation exposure pathways. A combination of pathways with a subset of routes might be involved depending on product use conditions and requirements, such as label instructions, actuation, and environmental conditions that influence the nature of the residues (Dary et al., 1996). The actual mass transfer to the membranes is dependent on contact with residues in air, in water, in food, and on surfaces. Occupational or incidental human contact with residues is a consequence of ergonomics and biomechanics of human activities in proximity to residues. Respiration of pesticides as gases (fumigants) and nasal-pharyngeal absorption and mucosal ingestion of pesticides as particles (soil/dust and particle type formulations) require contact to occur through occupational use or proximity to vapors or dispersion from sprays or foggers (Whalen et al., 2006). The extrathoracic region of the respiratory tract is impacted by gases, vapors, and particles ( 60 μm) that dissolve in the mucosal membrane of the nasopharynx, causing irritation (Gad, 2005) and mucus formation and consequent ingestion. Absorption by the upper respiratory tract (URT) mucosa would be modeled concurrently with bronchial and alveolar absorption with unabsorbed gases being exhaled as spurious biomarkers. Most PBTK models have, as part of model structure, exposure input algorithms that link exposure to absorption or enteral (intraperitoneal, sublingual, or gavage ingestion) and parenteral (intravenous infusion or intramuscular or subcutaneous injection) introduction of a test substance directly into the vascular compartment (Blancato et al., 2004; Pratt, 1990). Therapeutic drug applications (Bischoff and Brown, 1966; Bischoff et al., 1971; Collins et al., 1982; Farris et al., 1988) provide rare and extremely valuable human clinical data, PBPK parameters, and structures that bypass dermal, inhalation, and gastrointestinal (GI) absorption with direct introduction into the vascular compartment. Rapid and often instantaneous delivery of drug may be used to “deconvolute”

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FIGURE 44.10 Example of a PBTK model implemented in U.S. EPA’s Exposure Related Dose Estimating Model (ERDEM) system.

(Fisher et al., 1985; Paustenbach and Leung, 1993) the absorption process (U.S. EPA, 2009b). It is absorption that defines the exposure to the ADMET process (Darvas and Dormán, 2002) in a PBTK model. PBTK models commonly use a mass balance around a compartment, accounting for inflow, outflow, partitioning, and metabolism. The mass balance equation for the fat, slowly perfused, rapidly perfused, and kidney compartments is typically modeled as Vj

dc j dt

 Q j (carterial  cv, j )  R j

cv, j  Rj 

cj Pj : blood V max j  c j

(14)

Km j  c j Q j  QFj  Qcardiac V j  VFj  BW  D where cj is the concentration of the chemical in the jth compartment (with carterial denoting the concentration in the arterial blood), Qj is the blood flow to the jth compartment, Vj is the volume of the jth compartment, cv,j is the concentration in the blood exiting the jth compartment, and Vmaxj and Kmj are the Michaelis-Menten constants for the jth compartment.

44.3.4 Exposure Reconstruction and Inverse Modeling As mentioned in Chapter 45, biomarkers can be used for reconstructing exposures. Estimation of exposures from biomarker data involves inverse modeling (or model inversion); the corresponding forward model involves estimation of dose profiles (model outputs) given the exposure profile (model inputs) and the physiological and biochemical parameters (model parameters). Inversion of models can be very complex because such problems usually do not have a unique “answer,” so solution methods are generally concerned with finding the most likely solution, or a distribution with respect to predefined objective criteria. The (potential) role of PBPK modeling in exposure reconstruction is described in the U.S. EPA report on PBPK modeling (U.S. EPA, 2005). A variety of comprehensive monographs are available on inversion techniques and applications (e.g., Aster et al., 2005; Kaipio and Somersalo, 2005; Liu and Han, 2003; Ramm, 2005; Tarantola, 1987, 2005; Vogel, 2002), including related areas such as regularization (e.g., Engl et al., 1996; Morozov and Stessin, 1993) and optimization (e.g., Floudas and Pardalos, 1996; Gelb, 1974; Loose et al., 2004; Nocedal, 2006). These techniques can be generally categorized into deterministic and stochastic strategies (Moles et al., 2003). Classical deterministic strategies include nonlinear regression maximum likelihood estimation (MLE) (Georgopoulos et al.,

Chapter | 44

Modeling and Predicting Pesticide Exposures

1994; Roy et al., 1996): Given the model parameter set and data, MLE techniques aim to maximize the likelihood (conditional probability, given a set of parameter values) that differences between data and model results are due to random error. The MLE approach has been used to estimate metabolic parameter values from in vivo data but the approach has limitations with respect to simultaneously estimating multiple parameters or exposures, as all the parameters are not simultaneously identifiable (Smith and Evans, 1995). Stochastic methods such as the Bayesian Markov Chain Monte Carlo (MCMC) (Gelman, 2004) can provide probabilistic estimates of an arbitrarily large number of parameters based on prior parameter information in the form of probability distributions and experimental data and are being applied increasingly in PBPK modeling (Covington et al., 2006; Gelman, 2004; Yokley et al., 2006). Exposure reconstruction can be pursued through deterministic or stochastic methods (Aster et al., 2005; Tarantola, 2005; Vogel, 2002). These methods in general utilize a systematic exploration of the input space to identify the “global” minimum of the error metric and the corresponding values of inputs. Stochastic methods randomly sample the input space and do not necessarily guarantee convergence to the best solution to the inversion problem. Deterministic methods can, in principle, guarantee the best solution, though not necessarily in a finite time (Moles et al., 2003). In contrast, stochastic methods can provide a reasonable solution with relative efficiency; this solution is in practice often the best available for modest computation times (Moles et al., 2003). Furthermore, stochastic methods are usually quite simple to implement and use, especially for “black box” models. In the deterministic approach, the exposure reconstruction problem can be formulated as a global minimization problem that involves finding possible exposures x by minimizing a “cost function” J, based on observed biomarker data b’(ti) at each time ti (a total of Nmeas measurements), and a forward model m(x,ti), which, in practice, can be a deterministic PBTK model. Additionally, constraints can be included in the form of ● ● ●

bounds on possible exposures (xL x xU), equality constrains on the model (f (x,b,t)  0), and inequality constraints (g (x,b,t 0)).

N meas

∑ (b(ti )  m(x, ti ))T (b′(ti )  m(x, ti )) i1

(15)

[ least square min nimization] and N meas

J  L(m, x, b)   ∏ f x (b(ti ) | x, m (x, ti )) i1

ood estimation] [maximum likeliho

where the likelihood L is expressed as a product of likelihood of the data at each point fx, which depends on the assumptions regarding the distribution of “errors” (i.e., differences between observations and model estimates). The stochastic approach usually involves a Bayesian approach. If prior knowledge on exposure attributes is represented by pprior(x), the theoretical (“model”) knowledge of the relation between x and b is represented by pmodel(b|x), and the prior information about exposure is represented by pprior(x), then for a specific set of biomarker measurements b’, perror (b|m) is the probability of measuring b when the true value is m; i.e., it is the distribution of “measurement error” and not “model error.” p(x, b)

p(x ) ∫ perror (b | m) pmodel (m | x ) dm ∫ p(x ) ∫ perror (b | m) pmodel (m | x ) dm dx

(17)

The term pmodel (m|x) can be understood as a delta function centered on m(x) (when the model is deterministic without error). For a deterministic model with model error (or uncertainty), it represents the distribution of uncertainty in m(x). For a stochastic model, pmodel (m|x) represents the distribution of predictions that can be obtained for a specific (fixed) set of inputs x. When adequate PBTK models of pesticides are available, they provide quantitative estimates of pmodel (m|x), which can be used in reconstruction using a black box type inversion. Sometimes, running a large number of simulations using PBTK models can become computationally challenging. Likewise, complex PBTK modeling scenarios (e.g., mixtures of metals or pesticides with large half-lives) may need significantly more computational time for a single simulation. Thus fast equivalent operating models (FEOMs), which are approximations of original models (e.g., see Balakrishnan et al., 2003; Li et al., 2002; Wang et al., 2005), are useful in order to achieve reasonable computational performance. A schematic representation of the major steps involved in exposure reconstruction is shown in Figure 44.11.

44.3.5 Sensitivity and Uncertainty Analyses

Typical examples of J include J

1011

(16)

It must be noted that the steps in exposure modeling can be applied in a “nested manner” in order to characterize both the uncertainty and population variability. In order to characterize population variability for a given set of exposure modeling options, the seven steps can be run to estimate the population variability. However, in order to characterize the uncertainty in estimates of population distributions of exposures, these calculations have to be run multiple times by sampling the corresponding parameter distributions that represent the uncertainties. Because of the wide variety of behaviors exhibited by people and the wide range of pesticide levels in environmental media contacted, Monte Carlo simulation is typically

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Potential Exposures Distribution

Distribution of Exposures Consistent with Biomarker data

Improved Sampling

tia en Po t

E*3

PBPK Model or FEOM run with E*N input

B*M

E*2 E*1 PBPK Model or FEOM run with E*1 input

B*3 B*2 B*1 Biomarker data (NHEXAS, NHANES, etc)

Exposure Model

Comparison with Biomarker Data

E*M

lE xp

os

ur

es

(S am

pl

es

)

Optimization Algorithms

Small proportion of rejected (unused) model simulations

Companion Data (“Exposure Related”)

FIGURE 44.11 Schematic representation of the use of exposure and toxicokinetic models in assessing exposures from available biomonitoring data. The schematic shows the use of optimized sampling in order to reduce the time required for reconstructing exposures. (Source: Georgopoulos et al., 2009)

used in the higher-tier models to quantify ranges of exposure (variability) in a simulated population. In a one-stage Monte Carlo simulation, a model is often run thousands of times to simulate individuals from a population of interest. For each iteration, one or more input parameter values are drawn randomly from probability distributions, resulting in a cumulative distribution function (CDF) of population exposures. In two-stage Monte Carlo modeling, a family of CDFs is generated, where each one reflects a one-stage Monte Carlo simulation. The difference among the multiple one-stage Monte Carlo CDFs in a two-stage Monte Carlo run is that the input value parameters vary to reflect uncertainty in model inputs. Each CDF represents the uncertainty about the percentile of people in the simulated population who have a given exposure or dose. The family of CDF curves represents the uncertainty about the exposure or dose received by a given percentile of the modeled population (Furtaw, 2001). ORD’s SHEDS models (Burke et al., 2001; Xue et al., 2006) use both one-stage and two-stage Monte Carlo sampling to quantify both variability and uncertainty in model inputs and outputs. The case study in Section 44.4.1 illustrates the application of SHEDS to assess children’s exposure to CCAtreated wood (see Xue et al., 2006; Zartarian et al., 2006). Approximation of uncertainties in probabilistic models can be achieved through the use of surrogate models such as the stochastic response surface method (SRSM) (Isukapalli et al., 1998), which allows for efficient analysis of uncertainty propagation (Isukapalli and Georgopoulos, 2001). Furthermore, when the models used are coded in either Fortran or C, further efficiencies can be obtained by applying SRSM in combination with sensitivity analysis software (ADIFOR/ADIC) employing automated differentiation of

computer code, developed at Argonne National Laboratory (Bischof et al., 1996; Isukapalli and Georgopoulos, 2001; Isukapalli et al., 2000). Furthermore, the high dimensional model representation (HDMR) method (Li et al., 2001; Wang et al., 2003) allows for faster sensitivity analysis and subsequent development of simpler, fast, but accurate “substitute” models. The HDMR method provides a “global” understanding of which model variables are significant in a dynamic system and how they are interrelated within the system. This is critical for complex models with several parameters/inputs, since it is important to identify those with the greatest effect on the model outputs.

44.4 DISCUSSION: MODELS TO SUPPORT PESTICIDE REGULATION The manufacture, sale, and application of pesticides are highly regulated in the United States. Therefore, the types of pesticides that the general population is exposed to vary from time to time. Even though some of the pesticides either have been voluntarily recalled or are banned in the United States, prior applications in specific microenvironments (such as carpets) may lead to high levels of exposures to these pesticides. Furthermore, for some classes of pesticides, such as organophosphate pesticides, significant amounts of exposure and biomonitoring data are available – these are useful not only for evaluating exposure models, but also for evaluating methods for exposure reconstruction. The examples in this section deal with estimating exposures to pesticides that are either voluntarily recalled or banned.

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Modeling and Predicting Pesticide Exposures

44.4.1 Chromated Copper Arsenate (CCA) Exposures to Children Chromated copper arsenate (CCA) wood preservatives have been commonly used to protect wood from deterioration. Although CCA registrants voluntarily cancelled CCA-treated wood for residential use, children may potentially be exposed to arsenic (As) and chromium (Cr) residues on surfaces of existing treated wood structures and in surrounding soil. To conduct their CCA risk assessment, EPA’s Office of Pesticide Programs (OPP) requested that the Human Exposure and Atmospheric Sciences Division (HEASD) conduct a probabilistic exposure assessment for children who frequently contact CCA-treated playsets and decks. The goal of this research was to support OPP’s risk assessment and develop, apply, peer review, and publish the results from this exposure assessment. The population of interest was 1- to 6-year-old children in the United States who frequently contact CCA-treated wood residues and/or soil containing As or Cr from public and residential playsets and decks. HEASD’s Stochastic Human Exposure and Dose Simulation model for wood preservatives (SHEDS-Wood) was applied to simulate lifetime, intermediate-term, and short-term population exposures and absorbed dose via residue ingestion, dermal residue contact, soil ingestion, and dermal soil contact pathways. Residue ingestion via hand-to-mouth contact was the most significant exposure route for most scenarios. Hand washing can significantly reduce exposures. SHEDS-Wood estimates are typically consistent with, or within the range of, other CCA exposure models. The five most important variables were wood surface residue-to-skin transfer efficiency, wood surface residue levels, fraction of hand surface area mouthed, time spent on/around playsets, and frequency of hand washing. For this assessment, variability was much greater than uncertainty for predicted population dose estimates due to parameter uncertainty. Positive external peer reviews of the SHEDS-Wood methodology (2002) and the CCA exposure assessment (2003) by OPP’s Scientific Advisory Panel enabled OPP to use the model and application for regulatory purposes. Close collaborations with the program office through all aspects of this research led to the successful use of the SHEDS-Wood exposure model results in OPP’s risk assessment. Sensitivity and uncertainty analyses provided information to guide future measurement research. The final version of the SHEDS-Wood code is on the EPA website for others to use, for CCA or other wood preservative assessments. SHEDS-Wood could be modified to address similar issues (e.g., children’s exposure to tire crumb materials used in playgrounds). This application is important for several reasons. OPP used SHEDS results directly in their final CCA risk assessment. SHEDS-Wood is now a peer-reviewed modeling tool for the agency to enhance risk assessment/management

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decisions and prioritize data needs related to wood preservatives. The exposure assessment will help states advise the public how to minimize health risks from existing treated wood structures. Industry is using SHEDS to estimate exposures to CCA and other wood preservatives. This is one of the agency’s first probabilistic modeling assessments for regulatory purposes. Figure 44.12 shows the results from the application of a comprehensive exposure model for CCA (SHEDS-Wood) (Xue et al., 2006; Zartarian et al., 2006) to study the exposures to children from CCA-treated wooden playsets and decks. The advantage of the population-based modeling is highlighted in Figure 44.13, which shows the overall uncertainty in the estimates of population exposures to CCA. The advantage of predictive modeling is highlighted in Figure 44.14, which shows the impact of alternative mitigation strategies.

44.4.2 Assessing Chlorpyrifos Exposures and Doses The organophosphate (OP) pesticides have been almost completely banned in recent years. However, they are still important for exposure modeling since they continue to be found in the environment. This may result from two sources. First, chemically active pesticides can be persistent in certain environments, such as in low ultraviolet light and sequestered conditions (e.g., embedded in carpet twill). Another source is that banned pesticides are stockpiled and may continue to be used, illegally, after bans. Chlorpyrifos was one of the most highly applied OP pesticides. It provides an excellent example to evaluate the performance of a population-based model, MENTOR/ SHEDS-Pesticides, a physically based probabilistic model using the comprehensive individual field measurements collected in the Children’s-Post-Pesticide-Application-ExposureStudy (CPPAES). A population-based model requires a thorough evaluation process with actual field data to ensure the predictions are realistic. Because duplicate diet samples were not collected as part of CPPAES, SHEDS-Pesticides simulated dose profiles did not account for the dietary route. Because of the types of data available in CPPAES (Hore et al., 2005, 2006), this dataset was used to test MENTOR/ SHEDS-Pesticides as a deterministic model for individual exposure/dose estimates to try to evaluate the SHEDS algorithms. MENTOR/SHEDS-Pesticides was used to simulate the daily exposure and dose profiles for seven of the CPPAES children for a period of 10 days post a chlorpyrifos application. The MENTOR/SHEDS-Pesticides simulations of daily exposure and dose profiles were compared with the CPPAES field measurements obtained over a 10-day period. This was the first attempt to evaluate the components of a probabilistic aggregate pesticide exposure model using precisely measured data collected from individual children.

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FIGURE 44.12 Estimates of chromated copper arsenate (CCA) exposures to children as estimated by SHEDS-Wood. LADD, lifetime average daily dose.

FIGURE 44.13 Uncertainty in the estimates of population exposures to chromated copper arsenate (CCA) through a two-stage Monte Carlo application of SHEDS-Wood. CDF, cumulative distribution function.

A major improvement from a typical SHEDS application was the use of CPPAES child-specific and activity-specific point estimates as input variables during model simulations. These activity estimates were extracted from the CPPAES activity diaries and field measurements. In previous applications of SHEDS-Pesticides, sampling activity has been from default input probability distributions (Zartarian et al., 2000). In addition, the modified MENTOR/SHEDSPesticides used the individual CPPAES child-specific activity diaries instead of randomly selecting extant diaries from the EPA CHAD (McCurdy et al., 2000).

CPPAES child-specific microactivities. CPPAES childspecific microlevel activity data (i.e., hand-mouth and objectmouth frequencies as extracted from videotaped information) were specified for each individual simulation. An average hand-mouth and object-mouth frequency of 10  7 events/h (range 1.0–18) and 5  4 (range 1.4–15) , respectively, was calculated for the CPPAES children. CPPAES environmental measurements. The MENTOR/ SHEDS-Pesticides simulations included variables associated with environmental chlorpyrifos levels that were measured daily or periodically within the CPPAES treated

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Hand Washing & 90% Reduction FIGURE 44.14 Impact of different mitigation strategies on the levels of exposures to chromated copper arsenate (CCA), as estimated by the SHEDS-Wood model (Xue et al., 2006; Zartarian et al., 2006).

.00020 .00018 .00016 Average Exposure (mg/kg/day)

rooms for the indoor air and the indoor surfaces of each child’s home. Indoor air. The air measurement of chlorpyrifos collected in the field between sampling days 0 and 1 was designated as the measurement for day 0 in the MENTOR/ SHEDS, field 1 to 2 corresponded with day 1, and so forth. Based on whether or not a child was in a treated room or untreated room during each macroactivity period, MENTOR/SHEDS assigned a value for the chlorpyrifos air concentration level. Since indoor air samples were only collected from the treated rooms, untreated room air concentrations were defined according to a default SHEDS untreated/treated air concentration ratio (Hore et al., 2006). Exposure modules. MENTOR/SHEDS-Pesticides-simulated inhalation exposure estimates for the 2-week postapplication period were within a factor of 1.5 from the corresponding field estimates. Simulated dermal-hand exposure estimates, on the other hand, were found to be within a factor of 5.0 from the corresponding CPPAES values estimated using 2-week handrinse/handwipe samples (Figure 44.15). Comparing a daily average SHEDS hand loading value with an instantaneous measure of each child’s hand as sampled in the field on each sampling day probably contributed toward some of the discrepancy. Closely matching an actual 1-h field-sampling period with the corresponding simulated macroactivity period, however, indicated that a more precise set of input parameters could improve the capabilities of SHEDS-Pesticides to simulate the dermal exposures, particularly when simulations

.00014 .00012 .00010 .00008 .00006 .00004 .00002 0.00000

N=7

SHEDS SHEDS

CPPAES

FIGURE 44.15 Boxplots for average day 0–10 inhalation exposures for seven CPPAES children (MENTOR/SHEDS-Pesticides vs. CPPAES) (mg/kg/day).

used toy wipe loadings and child-specific TCme/ma values (Figure 44.16). Toys, which represent many porous or cumulative surfaces routinely contacted by a child, seem to be a better representation of the chlorpyrifos available for contact on the hand or other skin surface. A large variability was observed in the simulated dermal-hand exposure estimates using LWW surface loadings and SHEDS default TCme/ma values as derived from the jazzercise study (Ross et al., 1990). This is not surprising, since the jazzercise study transfer coefficients were based on measurements collected from adults following

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Min-Max

25%-75%

Median value

5.0E-4

Average Exposure (mg/kg)

4.0E-4

3.0E-4

2.0E-4

1.0E-4

0

N=7

LWW / Child TC Median= 4.1x10-5

LWW/ SHEDS TC Median= 6.9x10-5

Toy W ipe/ Child TC Median= 8.5x10-5

CPPAES Median= 1.7x10-5

FIGURE 44.16 Boxplots for average day 0–10 dermal exposures for seven CPPAES children (MENTOR/SHEDS-Pesticides [simulations with LWW surface loading with child-specific TCme/ma; LWW with default SHEDS TCme/ma; toy wipe with child-specific TCme/ma] vs. CPPAES) (mg/kg).

an exercise routine, which would not adequately represent children’s activities, as children can be considerably more versatile than adults (Hubal et al., 2000). The advantage of using CPPAES-derived transfer coefficients instead of using default MENTOR/SHEDS-Pesticides transfer coefficients was the fact that the coefficients were based on measurements collected from the CPPAES children. Dose module. CPPAES TCPy levels, both CR adjusted and non-CR adjusted, were underestimated by the SHEDS simulated estimates. This, however, was expected since SHEDS estimates only accounted for exposures via the inhalation, dermal, and nondietary ingestion routes, not accounting for the dietary route. Despite the limitation, model-simulated TCPy estimates were within a factor of 8 from the corresponding field-measured levels. A complete data set would be needed, inclusive of dietary information, to fully evaluate the dose SHEDS-Pesticides module.

CONCLUSION Screening-level multimedia contaminant fate and exposure models are useful to decision makers because these models provide an appropriate quantitative framework to evaluate our understanding of the complex interactions between

chemicals and the environment. The greatest challenge for multimedia models is to provide useful information without creating overwhelming demands for input data and producing outputs that cannot be evaluated. The multimedia modeler must struggle to avoid making a model that has more detail than can be accommodated by existing theory and data while also including sufficient fidelity to the real system to make reliable classifications about the source-todose relationships of environmental chemicals. More comprehensive models provide realistic descriptions of the underlying processes and are invaluable for performing detailed analysis. Thus, the evolution of models will allow for more realistic scenarios, especially regarding personal exposures to pesticides. The U.S. EPA’s CREM (U.S. EPA, 2008a) and the National Research Council (NRC, 2007) recommend various types of evaluation of models used in regulatory decision making. The pesticide exposure assessment models being used currently for EPA regulatory purposes have undergone external peer reviews (e.g., by EPA OPP’s FIFRA SAP), have been published in the peer-reviewed literature, and have been evaluated to some extent through model-to-model comparisons and/or real-world biomonitoring data (in cases where the exposure models are linked with dose estimation models). Although there are many similarities across higher tier pesticide exposure models, there

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can be differences in pathway-specific exposure algorithms (e.g., simulating a child’s hand-to-mouth ingestion for repeated mouthing events), in databases used to simulate a population (e.g., U.S. Census, CSFII, CHAD), in exposure-related algorithms (e.g., simulating an individual’s longitudinal activity pattern based on cross-sectional diary data), and in how pesticide concentrations in environmental media are obtained (e.g., measured field data or estimates from pesticide fate and transport models). Because of these differences, EPA’s OPP takes the approach of considering for regulatory purposes outputs from multiple pesticide exposure models that have undergone peer review and other model evaluation efforts. There are a number of research needs related to pesticide exposure models to address various types of uncertainties. These include data needs for key model inputs (e.g., longitudinal activity pattern data, spatial and temporal variability in pesticide levels contacted in various media, information to simulate cumulative exposures to multiple pollutants or mixtures), better understanding of more complex exposure algorithms (e.g., simulating children’s handto-mouth and object-to-mouth ingestion), and evaluation of model predictions (via linked exposure and dose models) against real-world biomonitoring data. As new and modified formulations of pesticides arrive in the marketplace, models will need to adapt, especially as these pesticides’ physical and chemical characteristics change. In addition, as application practices, including those of homeowners, change, so should the human exposure factors used in pesticide models.

ACKNOWLEDGMENTS This work was supported in part by the University Partnership Agreement, including the U.S. EPA-funded Center for Exposure and Risk Modeling (CERMEPAR827033), the Environmental Bioinformatics and Computational Toxicology Center (GAD R 832721-010), and the interagency agreement with the Department of Energy (IAG DW89930582). Additional support was provided by NIEHS Center for Environmenal Exposures and Disease at EOHSI (Grant No. P01E511256-01). Although this chapter has been cleared for publication by the U.S. EPA, the contents of this work are solely the responsibility of the authors and do not necessarily represent the official views of the funding agencies.

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Comprehensive Chemical Exposure Framework Using Person Oriented Modeling”. Prepared by The Lifeline Group for The Exposure Technical Implementation Panel, American Chemistry Council. Contract Number 1338. http://www.thelifelinegroup.org/LifeLine/ Documents/comprehensive_chemical_exposure_framework.pdf Ramm, A. G. (2005). “Inverse Problems”. Springer Science, New York. Reddy, M. B., Yang, R. S. H., Clewell, H. J. III, and Andersen, M. E. (2005). “Physiologically based Pharmacokinetic Modeling: Science and Applications”. John Wiley & Sons, Hoboken, N.J. Ross, J., Thongsinthusak, T., Fong, H., Margetich, S., and Krieger, R. (1990). Measuring potential dermal transfer of surface pesticide residue generated from indoor fogger use: An interim report. Chemosphere 20, 349–360. Rossman, L. A. (2000). “EPANET 2 Users Manual”. U.S. Environmental Protection Agency. Cincinnati, OH. EPA/600/R-00/057. Roy, A., Weisel, C. P., Ballo, M. A., and Georgopoulos, P. G. (1996). Studies of multiroute exposure/dose reconstruction using physiologically based pharmacokinetic models. Toxicol. Ind. Health 12(2), 153–163. Scire, J. S., Strimaitis, D. G., and Yamartino, R. J. (2000). “A User’s Guide for the CALPUFF Dispersion Model (Version 5.4)”. Earth Tech, Inc, Concord, MA. Smith, A. E., and Evans, J. S. (1995). Uncertainty in fitted estimates of apparent in-vivo metabolic constants for chloroform. Fundam. Appl. Toxicol. 25(1), 29–44. Smith, T. J. (1992). Occupational exposure and dose over time: limitations of cumulative exposure. Am. J. Ind. Med. 21(1), 35–51. Stallings, C., Tippett, J.A., Glen, G., and Smith, L. (2002). “CHAD User’s Guide-Extracting Human Activity Information from CHAD on the PC”. Written for USEPA National Exposure Research Laboratory by ManTech Environmental Technologies. http://www.epa.gov/ chadnet1/reports/CHAD_Manual.pdf Tao, S. S. H., and Bolger, P. M. (1999). Dietary arsenic intakes in the United States: FDA total Diet Study, September 1991-December 1996. Food Addit. Contam. 16(11), 465–472. Tarantola, A. (1987). “Inverse Problem Theory: Methods for Data Fitting and Model Parameter Estimation”. Elsevier, New York, NY. Tarantola, A. (2005). “Inverse Problem Theory and Methods for Model Parameter Estimation”. Society for Industrial and Applied Mathematics, Philadelphia, PA. Thomas, K. W., Pellizzari, E. D., and Berry, M. R. (1999). Populationbased dietary intakes and tap water concentrations for selected elements in the EPA region V National Human Exposure Assessment Survey (NHEXAS). J. Expo. Anal. Environ. Epidemiol. 9(5), 402–413. Tippett, K. S., Enns, C. W., and Moshfegh, A. J. (1999). Food consumption surveys in the US Department of Agriculture. Nutr. Today 34(1), 33–46. Traas, T. P., Stab, J. A., Kramer, P. R. G., Cofino, W. P., and Aldenberg, T. (1996). Modeling and risk assessment of tributyltin accumulation in the food web of a shallow freshwater lake. Environ. Sci. Technol. 30(4), 1227–1237. Turk, B. H., Grimsrud, D. T., Brown, J. T., Geisling-Sobotka, K. L., Harrison, J., and Prill, R. J. (1989). Commercial building ventilation rates and particle concentrations. ASHRAE Trans. 95(part 1), 422–433. USCB. (2009). US Census Bureau Population Information Website (http://www.census.gov/). US Census Bureau. U.S. EPA. (1997). “Exposure Factors Handbook”. USEPA Office of Research and Development. Washington, DC. http://www.epa.gov/ NCEA/pdfs/efh/front.pdf U.S. EPA. (1998). “User’s Guide for the AMS/EPA Regulatory ModelAERMOD”. Research Triangle Park, NC: USEPA Office of Air Quality Planning and Standards.

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U.S. EPA. (1999). “User Manual for the EPA Third-Generation Air Quality Modeling System (Models-3 Version 3.0)”. USEPA, Office of Research and Development. Washington D.C. EPA-600/R-99/055. U.S. EPA. (2001). “Draft Protocol for Measuring Children’s NonOccupational Exposure to Pesticides by all Relevant Pathways”. USEPA Office of Research and Development. Research Triangle Park, NC. EPA/600/R-03/026 U.S. EPA. (2003). “Multimedia, Multipathway, and Multireceptor Risk Assessment (3MRA) Modeling System. Volume I: Modeling System and Science”. U.S. Environmental Protection Agency. Research Triangle Park, NC. EPA530-D-03-001a. U.S. EPA. (2005). “Approaches for the Application of PhysiologicallyBased Pharmacokinetic Models and Supporting Data in Risk Assessment”. USEPA National Center for Environmental Assessment. EPA/600/R-05/043A (External Review Draft). http:// cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid  135427. (Note: this report is informally cited here for informational purposes only as it is still under external review) U.S. EPA. (2006a). “Approaches for the Application of Physiologically Based Pharmacokinetic (PBPK) Models and Supporting Data in Risk Assessment”. USEPA National Center for Environmental Assessment. Washington DC. EPA/600/R-05/043F. U.S. EPA. (2006b). NHEXAS-National Human Exposure Assessment Survey [January 2007]. Available from http://www.epa.gov/nerl/ research/nhexas/nhexas.htm. U.S. EPA. (2007). “Pesticide Inert Risk Assessment Tool”. U.S. Environmental Protection Agency, Office of Pollution, Prevention, and Toxics, Exposure Assessment Tools and Models. Washington, DC. http://www.epa.gov/oppt/exposure/pubs/pirat.htm U.S. EPA. (2008a). Council for Regulatory Environmental Modeling (website). US Environmental Protection Agency. http://cfpub.epa. gov/crem/ U.S. EPA. (2008b). ERD 3MRA Page. Multimedia, Multi-pathway, Multi-receptor Exposure and Risk Assessment (3MRA). Available from http://www.epa.gov/athens/research/projects/3mra. U.S. EPA. (2008c). Information Available from the Safe Drinking Water Information System. Available from http://www.epa.gov/safewater/ sdwisfed/sfed2.html. U.S. EPA. (2009a). Exposure Related Dose Estimating Model (ERDEM). Available from http://www.epa.gov/heasd/products/erdem/erdem.htm. U.S. EPA. (2009b). Physiologically-Based Pharmacokinetic (PBPK) Modeling Approaches for Interpreting Exposure, Biomarker and Pharmacokinetic Data in Risk Assessments (in preparation). Vogel, C. R. (2002). “Computational Methods for Inverse Problems, Frontiers in Applied Mathematics”. Society for Industrial and Applied Mathematics, Philadelphia. Walko, R. L., and Tremback, C. J. (2001). “RAMS-The Regional Atmospheric Modeling System, Version 4.3/4.4, Introduction to

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RAMS 4.3/4, 4”. ASTER Division, Mission Research Corporation, Fort Collins, CO. Walko, R. L., Tremback, C. J., and Bell, M. J. (1999). “HYPACT The Hybrid Particle and Concentration Transport Model, Version 1.1.0 User’s Guide”. ASTER Division, Mission Research Corporation, Fort Collins, CO. Wang, S. W., Georgopoulos, P. G., Li, G., and Rabitz, H. (2005). Characterizing uncertainties in human exposure modeling through the Random Sampling-High Dimensional Model Representation (RSHDMR) methodology. Int. J. Risk Assess. Manag. 5, 387–406. Wang, S. W., Georgopoulos, P. G., Li, G. Y., and Rabitz, H. (2003). Random sampling-high dimensional model representation (RSHDMR) with nonuniformly distributed variables: Application to an integrated multimedia/multipathway exposure and dose model for trichloroethylene. J. Phys. Chem. A 107(23), 4707–4716. Whalen, J. E., Foureman, G. L., and Vandenberg, J. J. (2006). Inhalation risk assessment at the Environmental Protection Agency. In “Inhalation Toxicology” (H. Salem and S. A. Katz, eds.), 2nd ed. Informa HealthCare. WHO (2004). “Harmonization Project Document No. 1; Part 2: IPCS Glossary of Key Exposure Assessment Terminology”. World Health Organization. Williams, P. R. D., and Paustenbach, D. J. (2002). Risk characterization. In “Human and Ecological Risk Assessment: Theory and Practice” (D. J. Paustenbach, ed.). John Wiley & Sons, New York. Xue, J., McCurdy, T., Spengler, J., and Ozkaynak, H. (2004). Understanding variability in time spent in selected locations for 7-12year old children. J. Expo. Anal. Environ. Epidemiol. 14(3), 222–233. Xue, J., Zartarian, V. G., Ozkaynak, H., Dang, W., Glen, G., Smith, L., and Stallings, C. (2006). A probabilistic arsenic exposure assessment for children who contact chromated copper arsenate (CCA)-treated playsets and decks, Part 2: Sensitivity and uncertainty analyses. Risk Anal. 26(2), 533–541. Yokley, K., Tran, H. T., Pekari, K., Rappaport, S., Riihimaki, V., Rothman, N., Waidyanatha, S., and Schlosser, P. M. (2006). Physiologicallybased pharmacokinetic modeling of benzene in humans: a Bayesian approach. Risk Anal. 26(4), 925–943. Zartarian, V., Bahadori, T., and McKone, T. (2005). Adoption of an official ISEA glossary. J. Expo. Anal. Environ. Epidemiol. 15(1), 1–5. Zartarian, V. G., Ozkaynak, H., Burke, J. M., Zufall, M. J., Rigas, M. L., and Furtaw, E. J. Jr. (2000). A modeling framework for estimating children’s residential exposure and dose to chlorpyrifos via dermal residue contact and nondietary ingestion. Environ. Health Perspect. 108(6), 505–514. Zartarian, V. G., Xue, J., Ozkaynak, H., Dang, W., Glen, G., Smith, L., and Stallings, C. (2006). A probabilistic arsenic exposure assessment for children who contact CCA-treated playsets and decks, Part 1: Model methodology, variability results, and model evaluation. Risk Anal. 26(2), 515–531.

Chapter 45

Biomonitoring: Uses and Considerations for Assessing Nonoccupational Human Exposure to Pesticides Jon R. Sobus1, Marsha K. Morgan1, Joachim D. Pleil1 and ����������������� Dana B. Barr2 1

U.S. Environmental Protection Agency, Research Triangle Park, North Carolina U.S. Department of Health and Human Services, Centers for Disease Control and Prevention, Chamblee, Georgia

2

45.1  Introduction Biomonitoring is an important tool that can be used to evaluate human exposure to pesticides by measuring the levels of pesticides, pesticide metabolites, or altered biological structures or functions in biological specimens or tissues (Barr et al., 2005b; Needham et al., 2005, 2007). These measurements in biological media, referred to as biomarkers, reflect human exposure to pesticides through all relevant routes, and can therefore be used to monitor aggregate and cumulative exposures. Aggregate pesticide exposure is defined as exposure to a single pesticide from all sources, across all routes and pathways (U.S. EPA, 2001a). Cumulative pesticide exposure is defined as exposure to multiple pesticides that can cause the same toxic effect via a common biochemical mechanism (U.S. EPA, 2001a). The complexity of aggregate and cumulative pesticide exposures often obscures the linkages between exposure measurements and potential human health effects. Therefore, biomonitoring offers a means to clarify these critical relationships. However, careful interpretation of biomonitoring data is necessary to accurately assess human exposure to pesticides and the associated human health risks.

The purpose of this chapter is to provide an overview of the state of the science for pesticide biomonitoring research. We first present the fundamental concepts and primary uses of biomonitoring, and then highlight the major criteria required for the selection and use of biomarkers in population-based exposure studies. Next we focus on factors that affect the use and interpretation of biomarkers of exposure for current-use pesticides. We conclude by identifying critical data gaps and research needs in the field of biomonitoring; the consideration of these factors in future studies will better inform assessments of exposure, dose, and risk.

45.2  Linking pesticide exposure to health effects The relationships between human exposure to pesticides and possible health outcomes can be described using an exposure-effect continuum (Angerer et al., 2006; Needham et al., 2005, 2007; NRC, 1987). As shown in Figure 45.1, major components of this continuum include exposure, internal dose, biologically effective dose, early ­ biological

Biomonitoring

Exposure

Internal Dose

Biologically Effective Dose

Early Biological Effects

(ID)

(BED)

(EBE)

Health Effects

Figure 45.1  Generalized exposure-effect continuum for pesticides. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

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effects, and ultimate health effects. Pesticide exposure refers to human contact with pesticides in environmental media (Zartarian et al., 2005). Sources of pesticide exposure include such media as dust, soil, air, water, and food, and routes of pesticide exposure include inhalation, ingestion, and dermal contact. (The course that a pesticide takes from exposure source to exposure route is the exposure pathway (Zartarian et al., 2005.) The amount of pesticide that enters the human body after crossing an exposure surface (e.g., skin, lung tissue, gastrointestinal tract) is referred to as the internal dose (ID) (U.S. EPA, 2001b), and the amount of absorbed pesticide that reaches the target sites where biochemical alterations or adverse effects occur is the biologically effective dose (BED) (U.S. EPA, 2001b). The BED leads to early biological effects (EBE), which are the structural and functional impairments within the body (resulting from pesticide exposures) that correlate with, and possibly predict, the ultimate health effects (NRC, 2006). It is often difficult to directly relate exposure measurements to observed human health effects, considering the many sources and routes of environmental pesticide exposure. The exposure-effect continuum for pesticides can presumably be clarified with an understanding of the underlying biological functions and processes. In Figure 45.1, ID, BED, and EBE are highlighted as the key links between exposure and effects and are therefore the focus of biomonitoring studies. Unfortunately, these key links (especially BED and EBE) often occur in inaccessible human tissues (e.g., liver, lung, and brain) and are therefore difficult to measure directly. However, surrogate biological measurements of ID, BED, and EBE, and factors governing these measures, can be made in readily available human fluids and tissues including blood, urine, saliva, semen, skin, breast milk, and expired air. These surrogate biological measurements are referred to as biomarkers of exposure, effect, and susceptibility (NRC, 1987). Biomarkers of exposure include measurements of pesticides, pesticide metabolites, and modified molecules or cells (e.g., DNA and protein adducts) in biological tissues/ fluids (e.g., blood) or excreta (e.g., urine). These biological measurements are directly related to the dose of a pesticide (ID and BED in Figure 45.1) and are a function of pesticide exposure. Biomarkers of effect include measurements of biochemical, physiological, or behavioral alterations that result as a consequence of pesticide exposure. Some examples of biomarkers of effect include biological measurements of endogenous and inflammatory responses, measurements of DNA, protein, cell, tissue, and organ damage/modification, and observations of tumors or cancer cell clusters. These biological measurements reflect EBE (Figure 45.1), but are often difficult to ascribe to a specific pesticide exposure event.

Hayes’ Handbook of Pesticide Toxicology

Biomarkers of susceptibility include measurements of an individual’s inherent ability to respond to pesticide exposures. These measurements include observations of molecular properties and functions, such as genetic polymorphisms and enzyme activities, that can affect the rates of pesticide absorption, distribution, metabolism, and elimination (ADME), along with an individual’s biochemical disposition toward disease progression or repair. Biomarkers of susceptibility are affected by a suite of exogenous and endogenous sources, and therefore may be difficult to link to a specific pesticide exposure event. Although not apparent in Figure 45.1, complex biochemical processes (e.g., enzymatic activity), physiology (e.g., blood flow rates), and thermodynamics (e.g., tissue: blood partition coefficients) link ID to BED and EBE. We illustrate some of these intricate relationships using examples of biomarkers of exposure, effect, and susceptibility in Figure 45.2. In this figure, squares represent biomarkers of exposure (which can be measured inside or outside the body), ovals represent biomarkers of effect, and a triangle represents biomarkers of susceptibility. The directional arrows in Figure 45.2 represent the potential interactions between biological measures and processes. Since the biomarkers of effect and susceptibility shown in this figure are often difficult to attribute to a specific pesticide exposure event, the illustrated biomarkers of exposure, which include parent pesticides, pesticide metabolites, and pesticide adducts, remain the focus of pesticide biomonitoring research, and are therefore the focus of this chapter.

45.3  The uses of biomonitoring Biomarkers of exposure from samples of human tissue, fluids, and excreta offer qualitative or quantitative evidence of pesticide exposure. These measurements are particularly useful in exposure research because they can highlight population-based exposure trends and improve estimates of pesticide exposure and dose.

45.3.1  Assessing Population-Based Exposure Trends Biomonitoring is commonly used as a surveillance tool to identify baseline exposure levels in a population, trends in exposure levels over time, and unique subpopulations with higher exposure levels. Multiple biomonitoring studies have been conducted in the United States and abroad to evaluate human exposure to pesticides; examples of these studies are listed in Table 45.1. (A more comprehensive list of biomonitoring studies can be found elsewhere (Bouvier et al., 2005)). Several of these studies have focused on exposure to specific pesticides within particular subpopulations, including pregnant women (e.g., the Center for the

Chapter | 45  Biomonitoring: Uses and Considerations

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Genetic Polymorphisms, Enzyme Activity

Endogenous Responses

Environmental Exposure

Repair or Health Effects Inflammatory Responses

DNA, Protein, Cell, Tissue, and Organ Damage/ Modification

Tumors, Cancer Cell Clusters

Parent Compound Metabolites Protein and DNA Adducts Inside Body Outside Body Excreted Parent Compound

Excreted Metabolites

Figure 45.2  Biomarkers of exposure, effect, and susceptibility as part of the exposure-effect continuum. Squares represent biomarkers of exposure; ovals represent biomarkers of effect; triangle represents biomarkers of susceptibility; arrows represent the potential interactions between biological measures and processes.

Health Assessment of Mothers and Children of Salinas [CHAMACOS] study (Castorina et al., 2003)) and children (e.g., Children’s Total Exposure to Persistent Pesticides and Other Persistent Organic Pollutants [CTEPP] study (Morgan et al., 2005, 2007, 2008; Wilson et al., 2004), the Minnesota Children’s Pesticide Exposure Study [MNCPES] (Quackenboss et al., 2000), and the German Environmental Survey for Children [GerES IV] (Becker et al., 2008)). However, as clearly shown in Table 45.1, the Centers for Disease Control and Prevention’s (CDC’s) ongoing National Health and Nutrition Examination Survey (NHANES) is the most comprehensive source of pesticide biomonitoring data, providing thousands of yearly measurements of individual biomarkers, stratified by age, sex, and race/ethnicity (CDC, 2003a). The National Report on Human Exposure to Environmental Chemicals (NER), a publication of the NHANES data, allows scientists and health officials to evaluate the specific pesticides to which the U.S. population is commonly exposed, to track trends

in exposure levels over time, and to set priorities on human exposure and human health research efforts (CDC, 2002, 2003b, 2005). Numerous biomarkers are measured in the ongoing NHANES study to assess human exposure to organochlorine (OC) insecticides, organophosphate (OP) insecticides, carbamate insecticides, pyrethroid insecticides, and a variety of herbicides. Many OC insecticides for which biomarkers are measured (e.g., hexachlorobenzene and dichlorodiphenyltrichloroethane [DDT]) are no longer in use, or have restricted use in the United States. However, because of their relatively high persistence in the environment and in the body, these biomarkers can still be measured in human specimens such as blood. Unlike the persistent OC insecticides, many current-use OP insecticides (e.g., chlorpyrifos and malathion), carbamate insecticides (e.g., carbofuran and propoxur), pyrethroid insecticides (e.g., permethrin and deltamethrin), and herbicides (e.g., 2,4-dichlorophenoxyacetic acid [2,4-D] and

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Table 45.1  Sample Populations and Measurements Included in Pesticide Biomonitoring Studies Location (n)

Subjects (age)

Exposure Meas.

Biomarkers

Matrix

References

NHANES III (1988–1994)

United States (1000)a

Adults (20–59 years)

Questionnaire

PCP; 2,4,5-TCP; 2,4,6-TCP; TCPy; 1-NAP; 2-NAP; IPP; CFP; 2,4-D; 2,4-DCP; 2,5-DCP; 4-NP

Urine (spot)

Hill et al. (1995)

NHANES (1999–2000)

United States (2000)a

Children and adults (6–59 years)

Questionnaire

PCP; 2,4,5-TCP; 2,4,6-TCP; TCPy; 1-NAP; 2-NAP; IPP; CFP; 2,4-D; 2,4-DCP; 2,5-DCP; HCB; beta-HCH; gamma-HCH; p,p’-DDT; p,p’-DDE; o,p’-DDT; oxychlordane; HCE; TNC; Mirex; DMP; DMTP; DMDTP; DEP; DETP; DEDTP; MDA; PNP; IMPy; 2,4,5-T; AZM; AM; DEET; OPP

Blood, urine (spot)

Barr et al. (2004, 2005a); CDC (2003b)

NHANES (2001–2002)

United States (2500)a

Children and adults (6–59 years)

Questionnaire

PCP; 2,4,5-TCP; 2,4,6-TCP; TCPy; 1-NAP; 2-NAP; IPP; CFP; 2,4-D; 2,4-DCP; 2,5-DCP; HCB; beta-HCH; gamma-HCH; p,p’-DDT; p,p’-DDE; o,p’-DDT; oxychlordane; HCE; TNC; Mirex; DMP; DMTP; DMDTP; DEP; DETP; DEDTP; MDA; PNP; IMPy; 2,4,5-T; AZM; AM; ACM; MM; DEET; OPP; Aldrin; Dieldrin; Endrin; DEAMPY; CMHC; FPBA; cis-DCCA; trans-DCCA; DBCA; 3-PBA

Blood, urine (spot)

CDC (2005)

GerES II (1990–1992)

Germany (1990)

Children and adults (6–14 years; 25–69 years)

Environmental

PCP

Urine (spot)

Seifert et al. (2000)

GerES III (1998)

Germany (2800)b

Adults (18–69 years)

Environmental

alpha-HCH; beta-HCH; gamma-HCH; 2,4-DCP; 2,5-DCP; 2,6-DCP; 2,3,4-TCP; 2,4,5-TCP; 2,4, 6-TCP; 2,3,4,6-TeCP; PCP; p,p’-DDE

Blood, urine (spot)

Becker et al. (2002, 2003)

GerES IV (2003–2006)

Germany (1000)c

Children (3–14 years)

Environmental, questionnaire

alpha-HCH; beta-HCH; gamma-HCH; 2,4-DCP; 2,5-DCP; 2,6-DCP; 2,3,4-TCP; 2,4,5-TCP; 2,4, 6-TCP; 2,3,4,6-TeCP; PCP; p,p’-DDE; HCB; DMP; DMTP; DMDTP; DEP; DETP; DEDTP; FPBA; cis-DCCA; trans-DCCA; DBCA; 3-PBA

Blood, urine (spot)

Becker et al. (2008)

NHEXAS (1995–1996)

Maryland (80)

10 years old

Personal, environmental, questionnaire

1-NAP; TCPy; MDA; AM

Urine (repeated)

MacIntosh et al. (1999)

NHEXAS (1995–1997)

Arizona (179)

5 years old

Personal, environmental, activities

1-NAP; TCPy; MDA; AM

Urine (spot)

Robertson et al. (1999)

MNCPES (1997)

Minnesota (102)

Children (3–12 years)

Personal, environmental, questionnaire, activities

2,4-D; TCPy; MDA; AM

Urine (repeated)

Quackenboss et al. (2000)

Hayes’ Handbook of Pesticide Toxicology

Study (years)

Washington (110)

Children (2–5 years)

Questionnaire

DMP; DMTP; DMDTP; DEP; DETP; DEDTP

Urine (spot)d

Lu et al. (2001)

Columbia (1999)

New York (314)

Pregnant women and their children (mothers: 18–35 years)

Personal, questionnaire

Chlorpyrifos; diazinon; IPP

Blood (repeated)

Whyatt et al. (2004)

CHAMACOS (1999)

California (460)

Pregnant women and their children (mothers: 18 years)

Environmental, pesticide use

DMP; DMTP; DMDTP; DEP; DETP; DEDTP

Urine (repeated)

Castorina et al. (2003)

Mt. Sinai (1998–2002)

New York (400)a

Pregnant women and their children

Questionnaire

PCP; TCPy; 3-PBA; PON status

Blood, urine (repeated)

Berkowitz et al. (2003, 2004)

CTEPP (2000–2001)

North Carolina and Ohio (257)e

Children and caregivers (children: 1.5–5 years)

Personal, environmental, questionnaire

2,4-D; TCPy; PCP; 3-PBA; IMP

Urine (repeated)

Morgan et al. (2005, 2007, 2008); Wilson et al. (2004)

CPES-WA (2003–2004)

Washington (23)

Children (3–11 years)

Questionnaire

FPBA; cis-DCCA; trans-DCCA; DBCA; 3-PBA; MDA; TCPy; IMPy; DEAMPY; CMHC

Urine (repeated)

Lu et al. 2006a,b)

Key: 1-NAP, 1-naphthol; 2-NAP, 2-naphthol; 2,3,4,6-TeCP, 2,3,4,6-tetrachlorophenol; 2,3,4-TCP, 2,3,4-trichlorophenol; 2,4,5-T, trichlorophenoxyacetic acid; 2,4,5-TCP, 2,4,5-trichlorophenol; 2,4,6-TCP, 2,4,6trichlorophenol; 2,4-D, 2,4-dichlorophenoxyacetic acid; 2,4-DCP, 2,4-dichlorophenol; 2,4-DCP, 2,4-dichlorophenol; 2,5-DCP, 2,5-dichlorophenol; 2,6-DCP, 2,6-dichlorophenol; 3-PBA, 3-pPhenoxybenzoic acid; 4-NP, 4-nitrophenol; ACM, acetochlor mercapturate; AM, alachlor mercapturate; AZM, atrazine mercapturate; beta-HCH, beta-hexachlorocyclohexane; CHAMACOS, Center for the Health Assessment of Mothers and Children of Salinas; CFP, carbofuranphenol; cis-DCCA, cis-3-(2,2-dichlorovinyl)-2,2,-dimethylcyclopropane carboxylic acid; CMHC, 3-chloro-7-hydroxy-4-methyl-2H-chromen-2-one/ol; CPES, Children’s Pesticide Exposure Study; CTEPP, Children’s Total Exposure to Persistent Pesticides and other Persistent Organic Pollutants; DBCA, cis-3-(2,2-dibromovinyl)-2,2,-dimethylcyclopropane carboxylic acid; DEAMPY, 2-(diethylamino)-6-methylpyrimidin-4-ol/one; DEDTP, diethyldithiophosphate; DEET, N,N-diethyl-3-methylbenzamide; DEP, diethylphosphate; DETP, diethylthiophosphate; DMDTP, dimethyldithiophosphate; DMP, dimethylphosphate; DMTP, dimethylthiophosphate; FPBA, 4-fluoro-3-phenoxybenzoic acid; gamma-HCH, gamma-hexachlorocyclohexane; GerES, German Environmental Survey; HCB, hexachlorobenzene; HCE, heptachlor epoxide; IMPy, 2-isopropyl-4methyl-6-hydroxypyrimidine; IPP, 2-isopropoxyphenol; MDA, malathion dicarboxylic acid; MM, metolachlor mercapturate; MNCPES, Minnesota Children’s Pesticide Exposure Survey; n, number of subjects; NHANES, National Health and Nutrition Examination Survey; NHEXAS, National Human Exposure Assessment Survey; o,p’-DDT, o,p’-dichlorodiphenyltrichloroethane; OPP, ortho-phenylphenol; p,p’-DDE, 1,1’-(2,2-dichloroethenylidene)-bis[4-chlorobenzene]; p,p’-DDT, p,p’-dichlorodiphenyltrichloroethane; PCP, pentachlorophenol; PNP, para-nitrophenol; PON, paraoxonase; TCPy, 3,5,6-tTrichloro-2-pyridinol; TNC, trans-nonachlor; trans-DCCA, trans-3-(2,2-dichlorovinyl)-2,2,-dimethylcyclopropane carboxylic acid.

Chapter | 45  Biomonitoring: Uses and Considerations

Urban Child Exposure Study (1998)

a

Approximate number of subjects. Blood measurements from approximately 2800 subjects; urine measurements from 692 subjects. c Blood measurements from approximately 1000 subjects; urine measurements from approximately 600 subjects. d Two seasonal spot samples collected (fall and spring). e 257 subjects includes only children and not caregivers. b

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atrazine) are environmentally and biologically nonpersistent (although some of their degradates may remain in the environment for a longer period of time). Therefore, biomarkers of these pesticides reflect more recent environmental exposures (i.e., hours or a few days). Since many of these nonpersistent pesticides are still in use, biomonitoring can be used to identify the current exposure trends, to aid in the design of mitigation strategies to reduce broadscale exposures, and to assess the effectiveness of exposure-mitigation efforts (CDC, 2005).

45.3.2  Improving Estimates of Exposure   and Dose In the absence of biomonitoring data, dose (i.e., ID and BED) can be approximated using deterministic or probabilistic exposure models coupled with classical toxicokinetic (TK) or physiologically based toxicokinetic (PBTK) ­models (a description of PBTK models is included in Chapter 44 of this handbook). Exposure models estimate route-specific exposures using measurements of contact duration and pesticide concentrations in environmental and personal samples (e.g., air, soil, dust, water, food, and skin samples). Since sample concentrations can vary over time and space (e.g., outside vs. inside), repeated measures are often necessary to highlight exposure variability and to improve exposure estimates. Additionally, since pesticide exposures can vary according to subject-specific traits and activities, exposure models must consider observations of behavioral (e.g., hand-to-mouth activities for nondietary ingestions and food intake for dietary ingestion) and physical (e.g., exposed skin surface area for dermal exposure and ventilation rates for inhalation exposure) sources of variability. Once route-specific exposure estimates are produced from the exposure models, they are used as the input terms for TK or PBTK models to generate a dose estimate. Numerical constants used in the kinetic models (describing the ADME processes) are frequently derived from in vitro measurements and from in vivo rodent studies. Therefore, uncertainty may be associated with these kinetic parameter estimates. Biomonitoring data can improve estimates of dose derived from exposure and kinetic models, since biomarker measurements consider all routes of pesticide exposure and all physical, behavioral, and physiological sources of variability. These data can also be used to evaluate and improve existing exposure and kinetic models that are needed in population exposure studies where biomonitoring data are not available. Furthermore, biomonitoring can inform exposure-biomarker relationships using a forward dosimetry approach and can be used to work backward from biomarker measurements to exposure estimates using a reverse dosimetry approach. This information is particularly useful to support the human health risk assessment of pesticides.

Hayes’ Handbook of Pesticide Toxicology

45.3.2.1  Forward Dosimetry Forward dosimetry is an approach that can be used to understand the quantitative relationships between pesticide exposures and observed biomarker concentrations. In forward dosimetry, estimated or measured pesticide concentrations from environmental and personal (nonbiological) samples are used as inputs into probabilistic or deterministic exposure models and kinetic models to estimate pesticide dose. The dose estimate (based on aggregate intake) is then compared to a measured biomarker concentration; the results from this comparison provide necessary information regarding the important sources and routes of human exposure to pesticides and can be used to identify missing sources and routes of exposure. This information is valuable for the interpretation of existing biomarker data and for the design and execution of future population-based exposure studies. Forward dosimetry can also be used to estimate biomarker levels resulting from pesticide exposures at regulatory/guidance levels (e.g., reference doses or concentrations [RfDs and RfCs]) that are considered to be acceptable or safe (Hays et al., 2007). Comparing these estimated values to observed levels from population-based biomonitoring studies is useful for the assessment of human health risks. The methods and applications of forward dosimetry are more thoroughly discussed in Chapter 44. To date, few exposure and biomonitoring studies have been designed to use this forward dosimetry approach (Morgan et al., 2005, 2007; Wilson et al., 2007).

45.3.2.2  Reverse Dosimetry Reverse dosimetry (exposure reconstruction) is an approach that can be used to work backward from biomarker measurements to estimates of human exposure to environmental pesticides. Reverse dosimetry, like forward dosimetry, requires the use of exposure models and kinetic models to address heterogeneity in environmental exposure measurements and in the rates of ADME. This method, utilizing modeling results and measurements of biomarkers from observational studies, has the potential to yield exposure estimates that can be compared to regulatory/guidance levels (Hays et al., 2007). Unfortunately, reverse dosimetry requires the use of numerical model inversion techniques and does not yield a unique solution but, rather, a range of potential exposure scenarios (Clewell et al., 2008). Thus, there is uncertainty associated with exposure reconstruction estimates. Reducing uncertainty in exposure estimates will rely on an improved understanding of likely exposure scenarios and the factors affecting variability in toxicokinetic properties. Specific inversion techniques used to reconstruct environmental exposures are discussed in Chapter 44 and a more extensive review of up-to-date methods for exposure

Chapter | 45  Biomonitoring: Uses and Considerations

r­ econstruction, as well as the mathematical fundamentals of a computational framework, is presented in Georgopoulos et al. (2009). To our knowledge, there are currently no observational studies of environmental pesticide exposure that have been designed to use this approach.

45.4  Biomarker selection and use Biomarkers of exposure should be, at a minimum, sensitive, specific, valid, biologically relevant, and easy to collect (i.e., practical) in order to be useful as a surveillance tool and for improving quantitative estimates of exposure and dose (Metcalf and Orloff, 2004). Here we examine against these criteria the most commonly used biomarkers of pesticide exposure. Table 45.2 lists the 45 individual biomarkers of pesticide exposure, grouped by pesticide class, that were measured in NHANES during 1999–2002 (CDC, 2005); these data represent the most comprehensive set of published pesticide biomarker data to date. Listed for each biomarker in this table is the sample matrix used for analysis (addressing the issue of practicality), the overall geometric mean (GM) published in CDC’s latest NER (addressing the issue of sensitivity), and the minimum number of parent compounds (addressing the issue of specificity). We use this information here to evaluate individual analytes as useful biomarkers of pesticide exposure.

45.4.1  Sensitivity Undetectable levels of pesticide biomarkers can be an indication of infrequent and/or low-level pesticide exposures or an indication of insufficiently sensitive analytical methods. It has been suggested that biomarkers should be measurable even at very low doses and should vary consistently and quantitatively with respect to exposure (Bernard, 1995; NRC, 1987). Despite advances in analytical techniques that have allowed the measurement of pesticides at ultra-trace levels, sensitivity issues still impair the quantitation of individual analytes in biological samples. In CDC’s most recent NER (CDC, 2005), GM values for individual biomarkers were not calculated when more than 40% of the biomarker measurements were below the analytical limit of detection (LOD). As shown in Table 45.2, out of a total of 45 individual pesticide biomarkers measured in NHANES during 1999–2002, a total GM value was calculated for only six biomarkers based on the most recent survey data (CDC, 2005). (Total GM values were reported for an additional seven biomarkers based on the 1999–2000 survey data; the apparent drop in sensitivity from the 1999–2000 data to the 2001–2002 data may indicate a decrease in exposure to individual pesticides over time.) Considering these data, there is insufficient sensitivity to measure the majority of these analytes as biomarkers of ­environmental

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pesticide exposure. We note that several computational methods can be used to impute values for measurements that fall below the analytical LOD (e.g., LOD divided by 2 or LOD divided by the square root of 2 (Hornung and Reed, 1990)). Discussions of the most appropriate ways to treat values below the analytical LOD have been published (Helsel, 2005; Pleil and Lorber, 2007); however, this is still a topic of much debate.

45.4.2  Specificity A useful pesticide biomarker should be specific for a parent compound of interest (Metcalf and Orloff, 2004). Specific pesticide biomarkers can be used to assess aggregate exposure, since the biomarker measurement reflects exposure to one parent compound from all exposure sources and through all exposure routes. Table 45.2 shows that over half (29 out of 45) of the biomarkers of pesticide exposure measured in NHANES during 1999–2002 were specific for a single parent compound (CDC, 2005) and, therefore, are suitable for the assessment of aggregate exposure. In many cases, measured pesticide metabolites are not specific biomarkers because they are common to multiple parent compounds. This situation is clearly demonstrated with the six dialkyl phosphate metabolites of OP insecticides, which include dimethylphosphate (DMP), dimethylthiophosphate (DMTP), dimethyldithiophosphate (DMDTP), diethylphosphate (DEP), diethylthiophosphate (DETP), and diethyldithiophosphate (DEDTP). These six metabolites can be produced from the metabolism of several different OP insecticides (e.g., chlorpyrifos, diazinon, malathion, and parathion). Therefore, when using these biomarkers to assess exposure, the relative contribution from each OP insecticide must be known to accurately quantify the contribution from a single parent compound. While nonspecific pesticide biomarkers are not ideal for assessing aggregate exposure, they can be useful for assessing cumulative exposure, which involves exposure to multiple parent compounds involving a common mechanism of toxicity. In a case study of the CHAMACOS cohort (Castorina et al., 2003), the six nonspecific dialkyl metabolites of OP insecticides (i.e., DMP, DMTP, DMDTP, DEP, DETP, and DEDTP) were measured in the urine of 446 pregnant women to assess cumulative OP insecticide exposures. Here, OP insecticide cumulative dose equivalents (calculated using the relative potency factor [RPF  the ratio of the toxic potency of a given chemical to that of an index chemical] of each relevant OP insecticide in the cumulative assessment group) were calculated using nonspecific biomarker measurements to assess exposure risks for the pregnant women (Castorina et al., 2003). This application demonstrates the utility of nonspecific biomarkers for the assessment of cumulative exposure and dose.

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Table 45.2  Characteristics of the Biomarkers of Pesticide Exposure (Grouped Here by Pesticide Family) Measured in NHANES during 1999–2002 Practicality

Sensitivity

Specificity

Pesticide family

Biomarker

Abbreviation

Matrix

GMa, b

# Parentse

Organochlorine insecticides

Hexachlorobenzene beta-Hexachlorocyclohexane gamma-Hexachlorocyclohexane Pentachlorophenol 2,4,5-Trichlorophenol 2,4,6-Trichlorophenol p,p’-Dichlorodiphenyltrichloroethane o,p’-Dichlorodiphenyltrichloroethane 1,1’-(2,2-dichloroethenylidene)-bis [4-chlorobenzene] Oxychlordane Heptachlor epoxide trans-Nonachlor Mirex Aldrin Dieldrin Endrin

HCB beta-HCH gamma-HCH PCP 2,4,5-TCP 2,4,6-TCP p,p’-DDT o,p’-DDT p,p’-DDE

Serum Serum Serum Urine Urine Urine Serum Serum Serum

nc ncc nc nc nc ncc nc nc 1.81

1 1 1 2 2 2 1 1 1

Oxychlordane HCE TNC Mirex Aldrin Dieldrin Endrin

Serum Serum Serum Serum Serum Serum Serum

0.070 nc 0.104 nc nc nc nc

1 1 1 1 1 2 1

Organophosphate insecticides

Dimethylphosphate Dimethylthiophosphate Dimethyldithiophosphate Diethylphosphate Diethylthiophosphate Diethyldithiophosphate Malathion dicarboxylic acid para-Nitrophenol 3,5,6-Trichloro-2-pyridinol 2-Isopropyl-4methyl-6-hydroxypyrimidine 2-(Diethylamino)-6-methylpyrimidin-4-ol/one 3-Chloro-7-hydroxy-4-methyl-2H-chromen-2-one/ol

DMP DMTP DMDTP DEP DETP DEDTP MDA PNP TCPy IMPy DEAMPY CMHC

Urine Urine Urine Urine Urine Urine Urine Urine Urine Urine Urine Urine

nc ncc nc ncc 0.457 nc ncd nc 1.76 nc nc nc

17 13 5 10 10 4 1 4 2 1 1 1

Pyrethroid insecticides

4-Fluoro-3-phenoxybenzoic acid cis-3-(2,2-Dichlorovinyl)-2,2-dimethylcyclopropane carboxylic acid trans-3-(2,2-Dichlorovinyl)-2,2dimethylcyclopropane carboxylic acid cis-3-(2,2-Dibromovinyl)-2,2-dimethylcyclopropane carboxylic acid 3-Phenoxybenzoic acid

FPBA cis-DCCA

Urine Urine

nc nc

1 3

trans-DCCA

Urine

nc

3

DBCA

Urine

nc

1

3-PBA

Urine

0.321

3

Carbamate insecticides

2-Isopropoxyphenol Carbofuranphenol

IPP CFP

Urine Urine

nc nc

1 4

Herbicides

2,4,5-Trichlorophenoxyacetic acid 2,4-Dichlorophenoxyacetic acid 2,4-Dichlorophenol Alachlor mercapturate Atrazine mercapturate Acetochlor mercapturate Metolachlor mercapturate

2,4,5-T 2,4-D 2,4-DCP AM AZM ACM MM

Urine Urine Urine Urine Urine Urine Urine

nc nc ncc ncd nc nc nc

1 1 1 1 1 1 1

Other pesticides

N,N-Diethyl-3-methylbenzamide ortho-Phenylphenol 2,5-Dichlorophenol

DEET OPP 2,5-DCP

Urine Urine Urine

nc ncc ncc

1 1 1

(CDC, 2005).nc  not calculated in most recent survey; percentage of results  LOD was too high. Total (all subjects) geometric mean from the most recent survey (either 1999–2000 or 2001–2002).

a

b

Serum concentration (ng/g); urine concentration (g/l). Total (all subjects) geometric mean calculated in 1999–2000 survey but not in 2001–2002 survey. d No data available for 2001–2002 survey. e Minimum number of parent compounds as identified in NER (not including environmental breakdown products). c

Chapter | 45  Biomonitoring: Uses and Considerations

45.4.3  Validity A selected biomarker of pesticide exposure should be a valid indicator of an underlying exposure event. That is, a biomarker measurement should accurately reflect the magnitude of exposure to a specific pesticide. Unfortunately, many pesticides break down in the environment to produce degradates that are chemically equivalent to biological metabolites. Therefore, biomarker levels can reflect exposure to the parent pesticides and to their environmental degradates. For example, the OP insecticide chlorpyrifos can degrade in the environment to 3,5,6-trichloro-2-pyridinol (TCPy), which is commonly measured as a urinary biomarker of chlorpyrifos exposure. In residential settings, exposures to chlorpyrifos and TCPy can occur from several sources such as soil, dust, air, and food and through several routes including inhalation, ingestion, and dermal contact (Morgan et al., 2005). This information, combined with the fact that toxicological research has shown that rats orally exposed to TCPy excreted all of it unchanged in their urine (Timchalk et al., 2007), indicates that humans likely excrete in their urine substantial amounts of unchanged TCPy as a function of environmental TCPy exposure. This scenario, particularly in residential settings, can lead to an overestimation of exposure to the parent compound when relying on biological metabolites as urinary biomarkers of exposure (Duggan et al., 2003). This issue is widely ­applicable in pesticide biomonitoring research, because any pesticide that is hydrolytically metabolized in the body is likely to be metabolized in the environment.

45.4.4  Biological Relevance A biomarker of exposure ideally should be relevant to the exposure-effect continuum shown in Figure 45.1 (Metcalf and Orloff, 2004; NRC, 2006). In other words, the most useful biomarkers not only reflect pesticide exposures, but also increase our knowledge of the underlying biological events that lead to potential health effects (Schulte and Talaska, 1995). We discussed in the previous section that TCPy is a commonly measured urinary metabolite of chlorpyrifos. Although not specific to chlorpyrifos, DEP and DETP can also be measured in the urine to assess environmental chlorpyrifos exposure. However, each of these urinary metabolites is the result of detoxifying biochemical reactions (Timchalk et al., 2007). Since the in vivo toxicity of chlorpyrifos is a result of bioactivation into chlorpyrifos-oxon (CPO), the measurement of CPO in a biological matrix is presumably more biologically relevant than that of a detoxified compound (Timchalk et al., 2007). Thus, provided that adequate analytical techniques exist, and the stability of the chemical in the matrix is sufficient, the measurement of a biologically relevant compound is preferable compared to the measurement of detoxification products. (Currently CPO is difficult

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to measure in samples of human blood (Timchalk et al., 2002)). However, by measuring products of both activation and detoxification, the underlying biological processes of the exposure-effect continuum may be better explained. Additionally, the measurement of multiple compounds can be useful in ­ identifying factors that may confound biomarker measurements (e.g., environmental metabolite residues and metabolic variations) (Timchalk et al., 2007).

45.4.5  Practicality To be useful in large-scale studies, biomarkers should be easy to obtain, store, and analyze (Metcalf and Orloff, 2004; NRC, 2006). As previously mentioned, biomarkers of exposure can be measured in samples of human tissues or fluids and in samples of human excreta. In large pesticide biomonitoring studies (see Table 45.1), blood and urine are the most commonly used human tissues/fluids and excreta, respectively, because they are, relative to other matrices (e.g., breast milk, adipose tissue, cord blood, feces), abundant in supply, collected using relatively noninvasive techniques (particularly urine), and can be analyzed using well-established methods. In fact, all of the NHANES biomarker measurements listed in Table 45.2 were derived from samples of human blood and urine (CDC, 2005). Still, appropriate quality assurance/quality control measures must be in place when using blood and urine for pesticide biomonitoring research. Biomarkers, whether parent compounds, metabolites, or adducts, may not be stable in a biological matrix (or in a sample preparation matrix [e.g., solvent]) if archived prior to analysis. Additionally, samples that are inappropriately collected or stored can be subject to chemical contamination. Any changes for which biomarker levels are not adjusted (i.e., loss due to instability or increase due to contamination) will influence chemical measurements and ultimately lead to inaccurate estimates of exposure and dose. Another major issue of practicality in biomonitoring studies is the collection of biological samples from sensitive subpopulations, such as children. Children may be more highly exposed to pesticides than adults, due to obvious differences in diet, environment, and daily activities. Increased exposure can have a particularly large impact on children considering their smaller body masses, immature physiological systems, and rapid physical development (Needham and Sexton, 2000; O’Rourke et al., 2000). Therefore, it is important to better understand children’s exposure to pesticides using biomonitoring. One of the main difficulties in using biomonitoring to assess pesticide exposures in children is the logistics of sample collection. For children who are able to provide blood or urine samples, it may be difficult to acquire the volume necessary for chemical analysis. In addition, urine collection for very young children may require alternative approaches such as using urine bonnets (collection devices placed under

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t­oilet seats), disposable diapers, or diapers with removable inserts (Hu et al., 2004).

45.5  Factors affecting the use and interpretation of biomarkers of exposure 45.5.1  Kinetics Toxicokinetic properties affect the concentrations of pesticides and pesticide metabolites in the body, the biological matrices best suited for biomarker analysis, and the variability in biomarker levels over time (Clewell et al., 2008; Hays et al., 2007). Thus, the selection, use, and interpretation of biomarkers should consider rate constants that describe ADME processes. Here we demonstrate, using simplified theoretical models, the impact of varying kinetic rate constants on predicted biomarker levels. For these examples, we use simple one-compartment toxicokinetic models and assume that kinetic processes are first order (i.e., dependent on the concentration of one reactant). Concentrations of pesticide metabolites in the blood are shown in response to a fictitious week-long random exposure profile (exposure during 5 days of the week but not during the weekend), repeated over 9 weeks. While these models do not consider nonlinear processes such as metabolic induction or saturation, they are suitable to demonstrate the importance of kinetic functions in biomonitoring studies. Elimination from the body is often discussed in terms of a pesticide’s biological persistence; this is typically measured as a biological half-life, which is the time required for the biomarker level to decrease by one-half in the absence of further input (i.e., exposure). Biologically persistent pesticides, including many OC insecticides, are eliminated from the body over the course of months or years, whereas biologically nonpersistent pesticides, including many current-use OP, pyrethroid, and carbamate insecticides and herbicides, are eliminated from the body over the course of hours or days (CDC, 2005). In Figure 45.3, we highlight the effect of elimination rate on pesticide levels in the body. Here we show the blood metabolite levels over time of three different pesticides, each with the same uptake rate, but with different elimination rates; biological half-lives in this example are postulated as 1 day, 1 week, or 4 weeks. The blood concentration curves in this figure demonstrate that the elimination rate linearly affects the eventual stable level of each biomarker. That is, doubling the biological half-life also doubles the eventual blood concentration. For the biomarker with a 4-week half-life (representing moderate biological persistence), levels in the blood vary in the short term in response to exposure, but also continue to increase over the entire 9-week period. In fact, the magnitude of the

Hayes’ Handbook of Pesticide Toxicology

short-term variability in relation to total variability is only marginal. Therefore, measurements of biomarkers of persistent pesticides that are made close in time are likely to be more similar than distant measurements. Considering this observation, biomarkers of persistent pesticides likely reflect exposures that have occurred over the previous months and even years, and can therefore be used to monitor seasonal effects and age-related effects in a population. For the pesticide with a 1-day half-life, levels in the blood also vary in the short term, but do not accumulate with time. Therefore, nonpersistent biomarkers reflect only the variability in recent exposures having occurred over the previous hours or days. These biomarkers are useful in evaluating recent exposure events, but require information regarding the time of sampling in relation to the time of exposure. While Figure 45.3 demonstrates that the elimination rate can affect the eventual stable biomarker levels, Figure 45.4 demonstrates that the uptake rate (including absorption, distribution, and metabolism) can also affect biomarker levels, albeit in a somewhat dissimilar fashion. In this example, the elimination rate is held constant with half-life of 1 week, and the uptake rate is varied; the uptake half-times in this example are postulated as 1 h, 4 h, or 24 h. Here we see the inverse effect of varying the elimination rate; that is, the faster the pesticide is absorbed, distributed, and metabolized, the higher the eventual steady state concentration. Additionally, the longer the uptake half-time, the smoother the short-term fluctuation in biomarker levels. This scenario is applicable when we consider the different routes of pesticide exposure. For example, exposure to a pesticide through inhalation, ingestion, and dermal contact can all contribute to the internal pesticide dose, and we can assume that the pesticide will be eliminated at the same rate regardless of the exposure route. However, uptake of the pesticide through the skin may occur more slowly than through the gut, and uptake through the gut may occur more slowly than through the lungs. Therefore, as shown in Figure 45.4, the magnitude and variability of the biomarker level will depend on route of exposure. These examples of the influence of rate constants on biomarker levels highlight the importance of using kinetic models when designing biomonitoring studies and evaluating existing biomarker data. Moreover, these examples highlight the importance of understanding not only the most recent environmental exposure, but also the exposure history for any given subject (i.e., between-subject differences in biomarker levels can occur because of true differences in environmental exposure levels or because of differences in recent exposure history). Finally, these examples highlight the importance of understanding the prominent routes of exposure when deciphering observed biomarker data. All of these factors are particularly important when using biomarker levels to quantify exposure and dose for the purpose of human health risk assessment.

Chapter | 45  Biomonitoring: Uses and Considerations

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Figure 45.3  Theoretical levels of pesticide metabolites in the blood over time resulting from a week-long random intermittent pesticide exposure (exposure during 5 days of the week but not during the weekend; repeated over 9 consecutive weeks), a constant rate of uptake (4-h half time), and an elimination half-life of 24 h, 1 week, or 4 weeks.

Figure 45.4  Theoretical levels of pesticide metabolites in the blood over time resulting from a week-long random intermittent pesticide exposure (exposure during 5 days of the week but not during the weekend; repeated over 9 consecutive weeks), a constant rate of elimination (1-week half-life), and an uptake half time of 24 h, 4 h, or 1 h.

45.5.2  Urinary Excretion Rate In the previous section, we highlighted the effects of kinetic rates on pesticide biomarker levels. We noted that, while biomarkers of persistent pesticides typically reflect

long-term exposure trends, those of nonpersistent biomarkers reflect recent exposure events. Moreover, we pointed out that the relative variability in closely spaced biomarker measurements of persistent pesticides is smaller than that of nonpersistent pesticides. Thus, biomarkers of nonpersistent

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pesticides are only useful in evaluating recent exposure events when information is known about the time of sampling in relation to the time of exposure. Further complicating the assessment of biomarkers of exposure to nonpersistent pesticides is the fact that most of these biomarkers are measured in the urine (see Table 45.2) and are therefore subject to variations in urinary excretion rates. Here we discuss the major considerations and current approaches for analyzing urinary biomarker measurements of current-use nonpersistent pesticides. Since pesticide exposures can vary over short periods of time, simple “snapshot” measurements of environmental and personal samples may not accurately assign the exposure profile. Likewise, since the concentrations of urinary biomarkers of nonpersistent pesticides track closely with recent exposures, snapshot measurements of urine concentration may also misclassify an exposure profile. Nevertheless, biomarker concentration measurements from “spot” urine samples are very common in observational studies (see Table 45.1) to minimize the cost of sample analysis and burden on study participants. First morning voids are often selected as representative spot urine samples because they integrate over the longest period of time; samples collected at some other random time during the day may also be employed primarily for convenience to the subject. Despite their widespread use, biomarker concentration measurements from spot urine

Hayes’ Handbook of Pesticide Toxicology

samples cannot provide a value for total excreted material, which is necessary to estimate previously absorbed dose. This would require some additional knowledge of total urine volume, elapsed time, and some measures of uptake and elimination kinetics. In Figure 45.5 we show respective levels of urinary biomarker concentration (ng/l) and total excreted biomarker mass (ng) over a 2-day period (here we assume a constant rate of urine production and allow reasonable variations in the time between excretions). Calculated levels of concentration and excreted mass vary in daytime and evening measurements as a result of randomly assigned episodic exposures of different durations and magnitudes. Of particular note in Figure 45.5 are the first morning void estimates where levels of excreted mass are considerably elevated, reflecting the longer period of accumulation in the bladder. This figure demonstrates that without adjusting for volume of the urine void and the time since last void, biomarker concentration levels can erroneously assign a pesticide dose. Our simple example in Figure 45.5 does not address additional real-world variability in rate of urine production due to interindividual differences in age, sex, ethnicity, etc., and intraindividual differences affected by hydration, work load, environmental conditions, and other factors. As such, a measurement of urinary concentration alone cannot be expected to provide anything more than a qualitative view into the current exposure state of the subject. Because

Figure 45.5  Respective levels of urinary biomarker concentration (ng/l; left y axis) and total excreted biomarker mass (ng; right y axis) over a 2-day period as a function of random episodic exposures of different durations and magnitudes. (Here we assume a constant rate of urine production and allow for variations in the time between excretions.)

Chapter | 45  Biomonitoring: Uses and Considerations

urine output is not constant either within measurements from an individual or among individuals, creatinine adjustment of urinary metabolite concentrations has typically been performed to correct for urine dilution among spot samples (Barber and Wallis, 1986). The thought process behind this procedure is if creatinine excretion is reasonably constant, then the presence of higher concentration of creatinine in the urine would indicate that the urine is concentrated (i.e., the person has a low hydration state prior to sample collection) and the presence of lower concentrations would indicate that the sample was dilute (i.e., the person had a high hydration state prior to sample collection); the variations in hydration state are adjusted for by dividing the urinary pesticide concentration by the creatinine concentration. Research has indicated, however, that for a given individual, metabolic variations including diurnal metabolism, disease status, vitamin supplementation, diet, and other factors can alter the rate of creatinine excretion (Boeniger et al., 1993). Furthermore, interperson variation in creatinine excretion is large, especially when the population is heterogeneous (i.e., multiple ages, sexes, race/ ethnicities) (Barr et al., 2005c). Thus, creatinine adjustment of urinary data in population studies will often introduce, rather than reduce, variability (Barr et al., 2005c). An alternative to creatinine adjustment is specific gravity adjustment (Hauser et al., 2004; Meeker et al., 2005). However, since specific gravity is largely dependent upon the dissolved solids in urine, which are primarily derived from creatinine, specific gravity and urinary creatinine are highly correlated. Urine samples can be collected over a period of time (e.g., 2-h, 8-h, 12-h, and 24-h urine voids) to better understand variations in urine volume and how they impact estimates of dose. Twenty-four-hour urine voids that are individually collected and assessed may provide the information necessary to calculate total excreted biomarker mass and to estimate dose. However, 24-h urine voids are often difficult to accurately collect, as they are burdensome to observational study participants. Additionally, 24-h urine voids require large amounts of storage space and many collection and analysis materials.

45.6  Summary and suggestions for future research The field of biomonitoring has greatly expanded over the past several decades (Needham et al., 2007). A wealth of biomarker data has been generated to better understand population-based pesticide exposures in the United States and abroad (see Table 45.1). Pesticide biomarker measurements allow the assessment of trends and variability in population-based exposures, the improvement of pesticide dose calculations, and the evaluation of existing exposure and dosimetric models. However, without additional information,

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such as exposure measurement and activity pattern data and modeling techniques, it is often difficult to determine the most important routes of exposure, the magnitude and duration of exposures, and the impact of these exposures on human health risks. While this additional information is sometimes lacking in large-scale biomonitoring studies (see Table 45.1), it is vital for improving our understanding of the exposure-effect continuum (see Figure 45.1). Forward dosimetry is an approach that can be used to relate exposure and biomonitoring data to better understand important sources and routes of pesticide exposure. Reverse dosimetry is an approach that can be used to infer pesticide exposures from biomonitoring data when exposure data are limited or unavailable. This application of biomonitoring is critically important for protecting human health because our laws, regulations, and guidelines are based on measurements of pesticides in environmental media. It is now possible, using these available approaches, to enhance our capability to interpret biomarker measurements as a function of aggregate and cumulative exposure. The proper application of these forward and reverse dosimetry approaches can improve our understanding of the relationships between exposure, dose, health effects, and risk. Biomarkers should be sensitive, specific, valid, biologically relevant, and practical to be useful in assessing exposure trends and in quantifying pesticide exposure and dose. In this chapter we have highlighted that many biomarkers of current-use pesticides may not be sufficiently sensitive to be useful biomarkers. Therefore, there is a need to improve our analytical capabilities and to develop new methods or alternative approaches to measure biomarkers of pesticide exposure in biological matrices; these new approaches should consider the measurement of biologically relevant analytes. Additionally, the increased use of specific biomarker measurements is warranted to better understand aggregate pesticide exposures. For nonspecific biomarkers, more exposure data are needed to determine the relative contribution from each parent compound. For biomarkers that are affected by environmental degradates in addition to biologic metabolites, there is a need to better characterize exposure to the parent compounds versus exposure to the environmental degradates. One of the biggest challenges facing pesticide biomonitoring research is the interpretation of measurements of nonpersistent current-use pesticides (i.e., OP, carbamate, and pyrethroid insecticides, and herbicides) in samples of human urine. Urinary measurements of nonpersistent pesticide biomarkers are difficult to interpret in terms of exposure and dose because of their highly variable nature (reflecting variable exposure levels) and other influential factors (e.g., rate of urine production and excretion and time of sample collection). Therefore, studies are needed to show that urinary biomarker levels can accurately reflect exposure and dose after making adjustments for the aforementioned sources of variability. Presumably, this should

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first be demonstrated using 24-h individual urine samples. Data can then be compared to values adjusted for creatinine concentration, specific gravity, or time of previous urination and volume of the current void. Comparisons between dose estimates fashioned from 24-h voids versus adjusted estimates will gauge the suitability of using refined measurement techniques to accurately decipher exposure and dose. Positive findings from these studies could be used to improve large-scale population-based studies, in which dosimetry approaches could be performed using appropriate and cost-effective techniques.

Conclusion Biomonitoring is an important tool that can be used to estimate pesticide exposure and dose, to evaluate existing exposure and dosimetric models, and to assess human health risks. In this chapter we discussed the primary uses of biomonitoring, highlighting applications in current studies, and introducing techniques that can be used to estimate pesticide exposure and dose (i.e., forward and reverse dosimetry). We then discussed, citing examples from recent studies, the importance of selecting pesticide biomarkers that are measured with adequate sensitivity, are specific to individual pesticides, are valid indicators of exposure events, are relevant to human health effects, and are practical to obtain, store, and analyze. Next, we introduced factors that can affect the interpretation of biomarker measurements; for measurements of current-use pesticides, these factors primarily include toxicokinetic properties and urinary excretion rates. We concluded this chapter by identifying critical data gaps and research needs in the field of biomonitoring.

Acknowledgments The United States Environmental Protection Agency through its Office of Research and Development funded and managed the research described here. It has been subjected to the Agency’s administrative review and approved for publication.

References Angerer, J., Bird, M. G., Burke, T. A., Doerrer, N. G., Needham, L., Robison, S. H., Sheldon, L., and Zenick, H. (2006). Strategic biomonitoring initiatives: moving the science forward. Toxicol. Sci. 93, 3–10. Barber, T. E., and Wallis, G. (1986). Correction of urinary mercury concentration by specific gravity, osmolality, and creatinine. J. Occup. Med. 28, 354–359. Barr, D. B., Allen, R., Olsson, A. O., Bravo, R., Caltabiano, L. M., Montesano, A., Nguyen, J., Udunka, S., Walden, D., Walker, R. D., Weerasekera, G., Whitehead, R. D. Jr., Schober, S. E., and Needham,

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L. L. (2005a). Concentrations of selective metabolites of organophosphorus pesticides in the United States population. Environ. Res. 99, 314–326. Barr, D. B., Bravo, R., Weerasekera, G., Caltabiano, L. M., Whitehead, R. D. Jr., Olsson, A. O., Caudill, S. P., Schober, S. E., Pirkle, J. L., Sampson, E. J., Jackson, R. J., and Needham, L. L. (2004). Concentrations of dialkyl phosphate metabolites of organophosphorus pesticides in the U.S. population. Environ. Health Perspect. 112, 186–200. Barr, D. B., Wang, R. Y., and Needham, L. L. (2005b). Biologic monitoring of exposure to environmental chemicals throughout the life stages: requirements and issues for consideration for the National Children’s Study. Environ. Health Perspect. 113, 1083–1091. Barr, D. B., Wilder, L. C., Caudill, S. P., Gonzalez, A. J., Needham, L. L., and Pirkle, J. L. (2005c). Urinary creatinine concentrations in the U.S. population: implications for urinary biologic monitoring measurements. Environ. Health Perspect. 113, 192–200. Becker, K., Kaus, S., Krause, C., Lepom, P., Schulz, C., Seiwert, M., and Seifert, B. (2002). German Environmental Survey 1998 (GerES III): environmental pollutants in blood of the German population. Int. J. Hyg. Environ. Health 205, 297–308. Becker, K., Mussig-Zufika, M., Conrad, A., Ludecke, A., Schulz, C., Seiwert, M., and Kolossa-Gehring, M. (2008). German Environmental Survey for Children 2003/06 (GerES IV): Levels of selected substances in blood and urine of children in Germany. Research Report 202 62 219, Federal Ministry of the Environment, Berlin, Germany. Becker, K., Schulz, C., Kaus, S., Seiwert, M., and Seifert, B. (2003). German Environmental Survey 1998 (GerES III): environmental pollutants in the urine of the German population. Int. J. Hyg. Environ. Health 206, 15–24. Berkowitz, G. S., Obel, J., Deych, E., Lapinski, R., Godbold, J., Liu, Z., Landrigan, P. J., and Wolff, M. S. (2003). Exposure to indoor pesticides during pregnancy in a multiethnic, urban cohort. Environ. Health Perspect. 111, 79–84. Berkowitz, G. S., Wetmur, J. G., Birman-Deych, E., Obel, J., Lapinski, R. H., Godbold, J. H., Holzman, I. R., and Wolff, M. S. (2004). In utero pesticide exposure, maternal paraoxonase activity, and head circumference. Environ. Health Perspect. 112, 388–391. Bernard, A. M. (1995). Biokinetics and stability aspects of biomarkers: recommendations for application in population studies. Toxicology 101, 65–71. Boeniger, M. F., Lowry, L. K., and Rosenberg, J. (1993). Interpretation of urine results used to assess chemical exposure with emphasis on creatinine adjustments: a review. Am. Ind. Hyg. Assoc. J. 54, 615–627. Bouvier, G., Seta, N., Vigouroux-Villard, A., Blanchard, O., and Momas, I. (2005). Insecticide urinary metabolites in nonoccupationally exposed populations. J. Toxicol. Environ. Health B Crit. Rev. 8, 485–512. Castorina, R., Bradman, A., McKone, T. E., Barr, D. B., Harnly, M. E., and Eskenazi, B. (2003). Cumulative organophosphate pesticide exposure and risk assessment among pregnant women living in an agricultural community: a case study from the CHAMACOS cohort. Environ. Health Perspect. 111, 1640–1648. Centers for Disease Control and Prevention (CDC). (2002). “National Report on Human Exposure to Environmental Chemicals.” Centers for Disease Control and Prevention (CDC). (2003a). “National Health and Nutrition Examination Survey.” Atlanta, GA. Available: http://www.cdc.gov/nchs/nhanes.htm Centers for Disease Control and Prevention (CDC). (2003b). “Second National Report on Human Exposure to Environmental Chemicals.”

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Biomonitoring: Uses and Considerations

Centers for Disease Control and Prevention (CDC), (2005). “Third National Report on Human Exposure to Environmental Chemicals.” Available: http://www.cdc.gov/exposurereport/report.htm Clewell, H. J., Tan, Y. M., Campbell, J. L., and Andersen, M. E. (2008). Quantitative interpretation of human biomonitoring data. Toxicol. Appl. Pharmacol. 231, 122–133. Duggan, A., Charnley, G., Chen, W., Chukwudebe, A., Hawk, R., Krieger, R. I., Ross, J., and Yarborough, C. (2003). Di-alkyl phosphate biomonitoring data: assessing cumulative exposure to organophosphate pesticides. Regul. Toxicol. Pharmacol. 37, 382–395. Georgopoulos, P. G., Sasso, A. F., Isukapalli, S. S., Lioy, P. J., Vallero, D. A., Okino, M., and Reiter, L. (2009). Reconstructing population exposures to environmental chemicals from biomarkers: challenges and opportunities. J. Expo. Sci. Environ. Epidemiol. 19, 149–171. Hauser, R., Meeker, J. D., Park, S., Silva, M. J., and Calafat, A. M. (2004). Temporal variability of urinary phthalate metabolite levels in men of reproductive age. Environ. Health Perspect. 112, 1734–1740. Hays, S. M., Becker, R. A., Leung, H. W., Aylward, L. L., and Pyatt, D. W. (2007). Biomonitoring equivalents: a screening approach for interpreting biomonitoring results from a public health risk perspective. Regul. Toxicol. Pharmacol. 47, 96–109. Helsel, D. R. (2005). More than obvious: better methods for interpreting nondetect data. Environ. Sci. Technol. 39, 419A–423A. Hill, R. H. Jr., Head, S. L., Baker, S., Gregg, M., Shealy, D. B., Bailey, S. L., Williams, C. C., Sampson, E. J., and Needham, L. L. (1995). Pesticide residues in urine of adults living in the United States: reference range concentrations. Environ. Res. 71, 99–108. Hornung, R. W., and Reed, L. D. (1990). Estimation of average concentration in the presence of nondetectable values. Appl. Occup. Environ. Hyg. 5, 46–51. Hu, Y., Beach, J., Raymer, J., and Gardner, M. (2004). Disposable diaper to collect urine samples from young children for pyrethroid pesticide studies. J. Expo. Anal. Environ. Epidemiol. 14, 378–384. Lu, C., Barr, D. B., Pearson, M., Bartell, S., and Bravo, R. (2006a). A longitudinal approach to assessing urban and suburban children’s exposure to pyrethroid pesticides. Environ. Health Perspect. 114, 1419–1423. Lu, C., Knutson, D. E., Fisker-Andersen, J., and Fenske, R. A. (2001). Biological monitoring survey of organophosphorus pesticide exposure among pre-school children in the Seattle metropolitan area. Environ. Health Perspect. 109, 299–303. Lu, C., Toepel, K., Irish, R., Fenske, R. A., Barr, D. B., and Bravo, R. (2006b). Organic diets significantly lower children’s dietary exposure to organophosphorus pesticides. Environ. Health Perspect. 114, 260–263. MacIntosh, D. L., Needham, L. L., Hammerstrom, K. A., and Ryan, P. B. (1999). A longitudinal investigation of selected pesticide metabolites in urine. J. Expo. Anal. Environ. Epidemiol. 9, 494–501. Meeker, J. D., Barr, D. B., Ryan, L., Herrick, R. F., Bennett, D. H., Bravo, R., and Hauser, R. (2005). Temporal variability of urinary levels of nonpersistent insecticides in adult men. J. Expo. Anal. Environ. Epidemiol. 15, 271–281. Metcalf, S. W., and Orloff, K. G. (2004). Biomarkers of exposure in community settings. J. Toxicol. Environ. Health A 67, 715–726. Morgan, M. K., Sheldon, L. S., Croghan, C. W., Jones, P. A., Chuang, J. C., and Wilson, N. K. (2007). An observational study of 127 preschool children at their homes and daycare centers in Ohio: environmental pathways to cis- and trans-permethrin exposure. Environ. Res. 104, 266–274. Morgan, M. K., Sheldon, L. S., Croghan, C. W., Jones, P. A., Robertson, G. L., Chuang, J. C., Wilson, N. K., and Lyu, C. W. (2005).

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Exposures of preschool children to chlorpyrifos and its degradation product 3,5,6-trichloro-2-pyridinol in their everyday environments. J. Expo. Anal. Environ. Epidemiol. 15, 297–309. Morgan, M. K., Sheldon, L. S., Thomas, K. W., Egeghy, P. P., Croghan, C. W., Jones, P. A., Chuang, J. C., and Wilson, N. K. (2008). Adult and children’s exposure to 2,4-D from multiple sources and pathways. J. Expo. Sci. Environ. Epidemiol. 18, 486–494. National Research Council (NRC). (1987). Biological markers in environmental health research. Committee on Biological Markers of the National Research Council. Environ. Health Perspect. 74, 3–9. National Research Council (NRC). (2006). Human Biomonitoring for Environmental Chemicals. The National Academies Press, Washington, DC. Needham, L. L., Calafat, A. M., and Barr, D. B. (2007). Uses and issues of biomonitoring. Int. J. Hyg. Environ. Health 210, 229–238. Needham, L. L., Ozkaynak, H., Whyatt, R. M., Barr, D. B., Wang, R. Y., Naeher, L., Akland, G., Bahadori, T., Bradman, A., Fortmann, R., Liu, L. J., Morandi, M., O’Rourke, M. K., Thomas, K., Quackenboss, J., Ryan, P. B., and Zartarian, V. (2005). Exposure assessment in the National Children’s Study: introduction. Environ. Health Perspect. 113, 1076–1082. Needham, L. L., and Sexton, K. (2000). Assessing children’s exposure to hazardous environmental chemicals: an overview of selected research challenges and complexities. J. Expo. Anal. Environ. Epidemiol. 10, 611–629. O’Rourke, M. K., Lizardi, P. S., Rogan, S. P., Freeman, N. C., Aguirre, A., and Saint, C. G. (2000). Pesticide exposure and creatinine variation among young children. J. Expo. Anal. Environ. Epidemiol. 10, 672–681. Pleil, J. D., and Lorber, M. N. (2007). Relative congener scaling of polychlorinated dibenzo-p-dioxins and dibenzofurans to estimate building fire contributions in air, surface wipes, and dust samples. Environ. Sci. Technol. 41, 7286–7293. Quackenboss, J. J., Pellizzari, E. D., Shubat, P., Whitmore, R. W., Adgate, J. L., Thomas, K. W., Freeman, N. C., Stroebel, C., Lioy, P. J., Clayton, A. C., and Sexton, K. (2000). Design strategy for assessing multi-pathway exposure for children: the Minnesota Children’s Pesticide Exposure Study (MNCPES). J. Expo. Anal. Environ. Epidemiol. 10, 145–158. Robertson, G. L., Lebowitz, M. D., O’Rourke, M. K., Gordon, S., and Moschandreas, D. (1999). The National Human Exposure Assessment Survey (NHEXAS) study in Arizona–introduction and preliminary results. J. Expo. Anal. Environ. Epidemiol. 9, 427–434. Schulte, P. A., and Talaska, G. (1995). Validity criteria for the use of biological markers of exposure to chemical agents in environmental epidemiology. Toxicology 101, 73–88. Seifert, B., Becker, K., Helm, D., Krause, C., Schulz, C., and Seiwert, M. (2000). The German Environmental Survey 1990/1992 (GerES II): reference concentrations of selected environmental pollutants in blood, urine, hair, house dust, drinking water and indoor air. J. Expo. Anal. Environ. Epidemiol. 10, 552–565. Timchalk, C., Busby, A., Campbell, J. A., Needham, L. L., and Barr, D. B. (2007). Comparative pharmacokinetics of the organophosphorus insecticide chlorpyrifos and its major metabolites diethylphosphate, diethylthiophosphate and 3,5,6-trichloro-2-pyridinol in the rat. Toxicology 237, 145–157. Timchalk, C., Nolan, R. J., Mendrala, A. L., Dittenber, D. A., Brzak, K. A., and Mattsson, J. L. (2002). A physiologically based pharmacokinetic and pharmacodynamic (PBPK/PD) model for the organophosphate insecticide chlorpyrifos in rats and humans. Toxicol. Sci. 66, 34–53.

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U.S. Environmental Protection Agency (U.S. EPA)., (2001a). “Draft Protocol for Measuring Children’s Non-Occupational Exposure to Pesticides by all Relevant Pathways.” EPA/600/R-03/026. Office of Research and Development, Research Triangle Park, NC. U.S. Environmental Protection Agency (U.S. EPA). (2001b). “Terms of Environment: Glossary, Abbreviations, and Acronyms.” EPA 175-B-97001. Communications, Education, and Public Affairs, Washington, DC. Whyatt, R. M., Rauh, V., Barr, D. B., Camann, D. E., Andrews, H. F., Garfinkel, R., Hoepner, L. A., Diaz, D., Dietrich, J., Reyes, A., Tang, D., Kinney, P. L., and Perera, F. P. (2004). Prenatal insecticide exposures and birth weight and length among an urban minority cohort. Environ. Health Perspect. 112, 1125–1132.

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Wilson, N. K., Chuang, J. C., Iachan, R., Lyu, C., Gordon, S. M., Morgan, M. K., Ozkaynak, H., and Sheldon, L. S. (2004). Design and sampling methodology for a large study of preschool children’s aggregate exposures to persistent organic pollutants in their everyday environments. J. Expo. Anal. Environ. Epidemiol. 14, 260–274. Wilson, N. K., Chuang, J. C., Morgan, M. K., Lordo, R. A., and Sheldon, L. S. (2007). An observational study of the potential exposures of preschool children to pentachlorophenol, bisphenol-A, and nonylphenol at home and daycare. Environ. Res. 103, 9–20. Zartarian, V., Bahadori, T., and McKone, T. (2005). Adoption of an official ISEA glossary. J. Expo. Anal. Environ. Epidemiol. 15, 1–5.

Chapter 46

Pesticide Exposure from Residential and Recreational Turf Dennis R. Klonne and J. Marshall Clark Toxicology & Exposure Assessment Services, Inc., Raleigh, North Carolina University of Massachusetts, Amherst, Massachusetts

46.1  Introduction Friday night high school football, pee-wee soccer, kids laying in the cool summer grass to watch the clouds, kneeling on the grass to trim around flower beds, or even playing golf are activities that have a common potential for exposure to pesticide residues from contact with treated turf. Although we often think of pesticide exposures as resulting from making applications to turf, ornamentals, and gardens in outdoor residential settings, there is also the potential for postapplication exposure through contact with the treated turf and foliage. The discussion here is limited to turf exposures, although exposures from those other foliage sources may also be of some concern in certain situations. Because many of the same activities can occur on both residential turf and recreational turf (e.g., parks and playfields), no distinction is made between them in this discussion – the differences are primarily one of scale. To gain an appreciation of the extent of pesticide applications in the residential home and garden market, we present the following data reported by the U.S. Environmental Protection Agency (EPA). In 1994 and 1995, 268 million pounds of active ingredients were applied by homeowners to the home and garden, which includes home lawns (U.S. EPA, 1999b). Another 150 million pounds of active ingredients were applied by professional applicators to residential homes and gardens. In 1994, 70 million of 95 million households used pesticides, and the average annual use per household was calculated to be 1.9 pounds of active ingredient. In 2008, the U.S. EPA reported that 67 million pounds of synthetic pesticides were used on lawns annually and $700 million was spent on those pesticides (U.S. EPA, 2008). Approximately 20 million acres are cultivated as residential lawns, and lawn care is a $25-billion-dollar-a-year industry. Thus, the evaluation of the residential turf sector for potential pesticide exposure is warranted. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

For an exposure to occur from turf, several factors are involved: (1) The turf must be treated, (2) the person must come into contact with the turf, and (3) the pesticide residue must transfer from the turf to the person. In this chapter, it is assumed that the turf is treated. Also, we are not concerned with how the pesticides were applied – only that pesticide residues exist on turf. Our discussion focuses on how contact is made with the turf and the transfer of residue from the turf to skin. Although inhalation exposure assessments are typically performed along with the dermal assessments for turf re-entry studies, personal review of many proprietary studies and the conclusion of the U.S. EPA (1997a) indicate that this route of exposure is very minimal in comparison to the dermal route. Residential and recreational turf is normally treated with herbicides, insecticides, and fungicides to maintain it in a via­ ble state of health. Residues can exist with varying half-lives depending on the stability of the active ingredient in sunlight and heat; bacterial degradation; adsorption to and absorption into the turf, thatch, and soil; precipitation or irrigation; and so on. Active ingredients can typically be found to degrade in a matter of hours to weeks (Harris and Solomon, 1992; Klonne et al., 2001; Rosenheck et al., 2001; U.S. EPA, 1999b; see Chapter 47). Thus, there is ample opportunity for people to be exposed to residues on turf, and that exposure will tend to be higher for a given activity the closer in time it is to the application process and for products that are more easily removed, or transferred, from the turf. Although the potential for exposure from treated turf has been acknowledged for many decades, the issue was not raised to a level warranting extensive field research until the advent of the Food Quality Protection Act (FQPA) of 1996. This legislation required age-specific exposure data and quantification of exposures from all residential sources and has spawned several large programs to assess residential pesticide exposures and human activities and the resultant models based on those data (Auyeung et al., 2006; Centers for Disease 1037

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Control and Prevention (CDC), 2005; � Driver et al., 2001;� Health and Environmental Sciences Institute, 2004; Hore et al., 2006; Needham et al., 2005; U.S. EPA, 1997a,b; Xue et al., 2007). Turf treated with pesticides was an obvious source of potential exposure. This realization resulted in a U.S. EPA data call-in for assessment of postapplication exposure from turf that spawned an industrywide task force and resulted in much new product-specific and generic data over nearly a decade (Outdoor Residential Exposure Task Force (ORETF), 2009; U.S. EPA, 1994, 1995). Even before the advent of FQPA, research was being done on postapplication exposure from turf. Although some studies were published, several others of which the authors are aware were proprietary studies done by pesticide companies in support of products with turf registrations. Thus, the early 1990s was a time when the methodology for these turf exposure assessment studies began to develop. It was not until the late 1990s that the U.S. EPA developed guidelines for conducting these studies (discussed later). Studies such as those briefly discussed later, demonstrated that exposure could occur from re-entering treated turf and provided the impetus for additional studies and refinement of methodologies. Adults who performed scripted activities (continuous 5-min cycles of walking, sitting, and lying on turf for a duration of 1 h) on turf at 1 h following a liquid spray application of 2,4-D at approximately 0.9 lb active ingredient (ai)/A were evaluated by biomonitoring (Harris and Solomon, 1992). Various levels of clothing were worn – pants, shortsleeved shirts, and shoes and socks, or shorts, short-sleeved shirts, and no footwear (although one person took off his shirt). Internal residues were not detected in fully clothed participants, whereas only three of five participants wearing shorts, short-sleeved shirts, and no footwear (including one with no shirt) had detectable residues. The average dose for the three participants with detectable residues was 3.1 g/kg. Transferable turf residue (TTR) was also measured using a cheesecloth-covered shoe shuffle technique (described later). Black (1993) measured the transfer of a fluorescent tracer sprayed on turf to adults performing a choreographed routine of children’s exercises (Mousercise) for 11 min and compared it to that of children (3–9 years old) playing randomly on such turf for 30 min. TTRs were assessed by wiping the turf (100 cm2) 10 times with a water-dampened gauze cloth. Tracer deposition on the skin was quantified using video imaging and found to be 148 g/h. Adults who participated in 4 h of activities (60 min each of Frisbee, touch football, and picnicking on a blanket, and 30 min each of sunbathing on a blanket and weeding) following a liquid application of chlorpyrifos at 4 lb ai/A had internal doses of approximately 7 g/kg as determined by bio­monitoring (Vaccaro et al., 1996). Participants wore tshirts and shorts and were barefoot for all activities with the exception of touch football. Turf was sampled by a drag sled method (discussed later).

Hayes’ Handbook of Pesticide Toxicology

Although performed after the passage of FQPA, the study by Bernard et al. (2001) provided data comparing two methods of assessing exposure. Adults who performed a Jazzercise routine on turf were evaluated by either passive dosimetry (cotton gloves, socks, and long underwear) or biomonitoring (while wearing minimal clothing) following a liquid spray application of chlorpyrifos at 1 lb/A. TTR was evaluated using the California roller technique (described later). The calculated dose from passive dosimetry averaged 0.7 or 2 g/kg (depending on which of two reported dermal absorption values were used), whereas that from biomonitoring was 1.3 g/kg. Thus, both exposure methods provided very similar estimates of exposure. Although the activities on golf course turf are distinctly different from those typically occurring on residential turf, potential exposures have been the subject of several studies. Golf course exposures are discussed in detail in Chapter 47, but one study is briefly discussed here because of the widespread occurrence of this type of recreational turf and to provide an idea of the comparative exposure to the studies discussed previously. Putnam and Clark (2007) studied potential exposure to golfers on turfgrass treated with chlorpyrifos (4 lb/acre) that was irrigated immediately after application. Exposure to individuals simulating 18 holes of golf during a 4-h interval was determined by either whole-body dosimetry or biomonitoring. Re-entry was 1 h after the conclusion of irrigation. Participants in the biomonitoring group wore short-sleeved shirts, shorts, ankle socks, golf caps, and golf shoes. Air concentrations were evaluated via personal air samplers, and TTR was assessed using the California roller method and dampened cheesecloth wipes. Average dermal chlorpyrifos exposures determined by whole-body dosimetry were calculated to be 0.13 or 0.42 g/kg (depending on which of two reported dermal absorption values were used), and total exposures obtained by adding calculated inhalation exposure (0.18 g/kg) were 0.31 or 0.60 g/kg. The absorbed chlorpyrifos dose determined by biomonitoring was 1.06 g/kg. These data indicate that both methodologies gave similar results. Determination of exposure potential from turf is not an easy exercise. It is not similar to monitoring exposure to agricultural workers and is much more difficult to perform. In an agricultural work setting, the activity (e.g., harvesting cabbage) is fairly constant and the surface area of the leaf foliage for sampling (typically with leaf punches) for determination of the dislodgeable residue is large. In contrast, the activities on residential turf are quite varied, the duration and frequency for those myriad activities are quite variable, and the activities are affected by many different variables such as the age of the person engaged in them (e.g., youngsters typically wrestle and roll around on turf, whereas adults do not; teenagers and adults cut grass, whereas toddlers do not). The intensity of the contact with the turf and the proximity to the time of application will also affect

Chapter | 46  Pesticide Exposure from Residential and Recreational Turf

exposure. Furthermore, the surface area of the leaf is small, thus prohibiting the use of leaf punches and requiring other techniques to determine transferable residues. The assessment of exposures from treated turf is a timeand capital-intensive endeavor, especially when data are developed for support of a registration that has requirements that can easily double the study cost. For example, to plan and conduct a 15-person study performing a set routine conducted according to Good Laboratory Practices with simultaneous TTR samples will probably take more than 1 year and cost approximately $1 million. In addition, such a protocol being submitted to support a pesticide registration will not only be reviewed by the U.S. EPA but also be reviewed for both ethics and science content by the Human Studies Review Board in compliance with a regulation that establishes requirements for the protection of subjects in human research (U.S. EPA, 2006).

46.2  Assessment of transferable   turf residues The general classes of pesticides used on turf include herbicides, fungicides, and insecticides (see Chapter 47). The chemicals involved in potential exposure from treated turf vary due to several factors, such as the turf type (coolseason grasses such as fescue versus warm-season grasses such as Bermuda), time of year (pre-emergent herbicides used in spring versus fungicides used in summer), and pest pressure (e.g., grubs, brown patch, and crabgrass). Different pesticide formulations transfer from turf at different rates, even with the same active ingredient in them. Transferability from turf to people is influenced by a multitude of factors, including the general formulation type (e.g., liquid versus granular), water solubility of the active ingredient, environmental stability of the active ingredient, and additives to the formulation (e.g., stickers, stabilizers, surfactants). The practice of watering in a product following an application (i.e., postapplication irrigation; see Chapter 47; Putnam and Clark, 2007) can also have a dramatic effect on the transferable residue. Although it usually decreases the transferable residues from products applied as liquid sprays, it can increase the residues from products applied as granules because it releases the active ingredient from the carrier (e.g., clay) on which it is coated. Some examples of this are provided in Table 46.1. This makes it very difficult to predict the transferability of the active ingredient in a product, and TTR data are considered to be product specific. Therefore, every pesticide formulation must have its own TTR study as per the 1995 U.S. EPA data call-in notice (U.S. EPA, 1995). The issue then becomes one of selecting the best method to determine TTR from the multitude of methods that are currently available. If one considers the structure of turf, it is made up of individual blades of grass that grow at various heights and

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densities that change over the seasons. The grass type can vary greatly from coarse-bladed St. Augustine grass to finebladed bluegrass. It is apparent that harvesting very small sections of turf to determine transferable residues has the potential to produce highly variable results. Many methods have been proposed to assess TTR (Cowell et al., 1993; Fortune, 1997; Klonne et al., 2001; Lewis et al., 1993; Rosenheck et al., 2001; U.S. EPA, 1997a, 1998). Some of the methods were actually used to determine transferable residues from indoor flooring surfaces and were adapted to use on turf (Lewis et al., 1993; Ross et al., 1990, 1991). One method involves collecting turf samples with a reel mower so that measured subsamples can be treated with a mild soap solution to determine the residues that are able to be removed from the grass surface (Hurto and Prinster, 1993). That is, the procedure takes care not to further rupture the grass blades to release any absorbed residues. Concurrently, grass plugs are collected to make surface area determinations by putting them onto tape and measuring them with a leaf area meter while leaf weight is also determined using similar samples collected from the turf. It is obvious that this method is very tedious and time-consuming. Another method involves the use of a weighted metal pan (32 lb to simulate the weight of a toddler) that has denim attached to the surface in contact with the turf and which is dragged for a set length (Lewis et al., 1993). The residues transferred to the denim are subsequently determined. This method is a variant of the original method developed by Vaccaro et al. (1996) that employed a denimcovered wood block weighted to provide pressure comparable to a 1-year-old child. A method in which adults attach a metal pan to their feet that has cheesecloth on the bottom to collect transferable residues from the turf has also been used (Thompson et al., 1984). Adults shuffle their feet for a set length, keeping the pan in contact with the turf. The residues transferred to the cheesecloth are subsequently determined. A roller technique has also been used in which a cylinder attached to a handle and covered with polyurethane foam that is in contact with the turf is rolled for a designated length (Lewis et al., 1994). This procedure is repeated over several unsampled turf patches for a set number of times. The residues transferred to the polyurethane foam are then determined. A method that has been adopted by much of the industry involved with pesticide production and exposure assessment (e.g., ORETF) is the California roller technique of Ross et al. (1990, 1991), which was originally used to determine residues on indoor carpeting. Comparison of the methods described previously, and some others, found this procedure to provide a good combination of sensitivity to residues, repeatability within subjects, reproducibility among subjects, and ease of use (Fortune, 1997; Klonne et al., 2001; Rosenheck et al., 2001). Briefly, the method employs

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Table 46.1  Transferable Residues After Sprays Have Dried and Dusts Have Settled Chemical

Transferable residue (g/cm2) as estimated by residential SOP (20% transferable)

Transferable residue (g/cm2) from empirical measurements

Method

Chlorpyrifos, spray (1 lb ai/acre)

2.24

0.03 (0.27%)

Foliar wash

Chlorpyrifos, granular (2 lb ai/acre)

4.48

0.139 (0.6%)

Drag sled

Chlorpyrifos, spray (2 lb ai/acre)

9

0.03 (0.06%)

Drag sled

Diazinon, granular (4 lb ai/acre)

9

0.13 dry (0.3%) 0.27 watered-in (0.63%)

Foliar wash

DDVP, spray (2 lb ai/acre)

4.48

0.1 (0.45%)

Foliar wash

Fonofos, granular (4 lb ai/acre)

9

0.006–0.11 dry (0.013–0.24%) 0.018–0.074 watered-in (0.039–0.16%)

Foliar wash

Malathion, spray (5 lb ai/acre)

12.2

0.57 (1%)

Cloth roller

Isofenfos, spray (flowable) (2 lb ai/acre)

4.48

0.91 (4%)

Foliar wash

Dithiopyr, microencapsulated (1 lb ai/acre)

2.24

0.36 (3.19%)

Foliar wash (20% acetonitrile)

Amidochlor, spray (4 lb ai/acre)

9

0.22 (0.4%)

Shoe

2,4-D 2-EHE, spray (1.7 lb ae/acre)

3.81

0.025 (1.46%)

Cloth roller

MCPA 2-EHE (1.544 lb ae/acre)

3.46

0.015 (0.845%)

Cloth roller

MCPA DMA (1.547 lb ae/acre)

3.46

0.07 (0.4%)

Cloth roller

2,4-DP-p DMA (0.596 lb ae/acre)

1.34

0.016 (1.2%)

Cloth roller

MCPP-p DMA (0.599 lb ae/acre)

1.34

0.13 (1.91%)

Cloth roller

2,4-DP-p 2-EHE (0.612 lb ae/acre)

1.37

0.013 (0.186%)

Cloth roller

2,4-D DMA (1.585 lb ae/acre)

3.55

0.13 (0.73%)

Cloth roller

Amidochlor, spray (4 lb ai/acre)

9

0.17 (0.37%)

PUF roller

ae, acid equivalent; ai, active ingredient; SOP, standard operating procedure. Reproduced in part from U.S. EPA (1999b, p. 24).

a 4-in.-diameter, 24-in.-long cylinder filled with material (e.g., lead shot or sand) attached to a handle as shown in Figure 46.1 (Fuller et al., 2001). The cylinder is rolled over a set length of cotton sheet and the residues transferred to the sheet are determined. The method is useful for both liquid and granular formulations. Additional information has been obtained to demonstrate the performance of the California roller. In an independent round-robin testing protocol, the U.S. EPA found that the California roller performed more reproducibly and was easier to use than the polyurethane foam (PUF) roller or the drag sled (Fortune, 1997). Subsequently, Williams et al. (2003) also demonstrated that the California roller was quite “rugged” in that neither weight of the roller nor number of rolls tended to have a significant effect on TTR. A variant of this roller was used to further demonstrate that TTR changed only approximately 2-fold when the roller

weight was increased by 10-fold to simulate the point pressures exerted on a surface by the full range of body weights for children through adults (Williams et al., 2008). Guidelines for the conduct of the TTR study have been provided by the U.S. EPA (1998). The guidelines recommend conducting the study at the highest labeled application rate, and sampling typically starts as soon as the application sprays are dry or the dusts have settled (for liquid and granular formulations, respectively). Samples are collected from the day of application (day 0) and typically for 72 h (or at least three half-lives for more persistent chemicals). A decline curve is fit to the data and the day 0 residues are used for exposure assessments because it is assumed that human postapplication contact could occur soon after the application is completed. A homogeneous application of a product to turf can be calculated to result in a certain amount of active ingredient

Chapter | 46  Pesticide Exposure from Residential and Recreational Turf

Front View

Handle Length - 48 in. (122 cm)

Side View

Metal Tubing Handle

PUF Foam or Pipr Insulation

PVC Pipe

Roller Width - 24 in. (61cm)

Roller Diameter 4 in. (10 cm)

Figure 46.1  Modified California roller for collecting transferable turf residues (reproduced from Fuller et al., 2001).

per unit area of surface (assuming a flat surface; i.e., 11.2 g ai per square centimeter of surface area per pound of applied active ingredient per acre). The process results from a TTR technique such as the California roller can be calculated in the same units (or simply as g ai/cm2 of surface area if the application rate is not important to subsequent calculations). This allows for the calculation of a transferable residue as a percentage of the application. Although results can vary widely, for making initial estimates in exposure assessments, personal review of many proprietary studies indicates that values of 1 and 0.1% are a reasonable approximation for percentage transferable residue for liquid and granular formulations, respectively. However, the U.S. EPA uses a somewhat higher default of 5% transferable residue regardless of formulation for initial tier 1 assessments (discussed later).

46.3  Activities on residential and recreational turf The activities that occur on turf are quite variable. They are related to many factors, such as the season, geography, and age of the person. The exposure to a person re-entering treated turf is related to these factors and others, such as proximity of re-entry time to application and frequency, duration, and intensity of the activity. Given the nearly infinite number of combinations of all these factors, it is apparent that this aspect of performing a postapplication exposure assessment is extremely complex and subject to many different approaches. Efforts have been made to determine the activities that people engage in on turf. One large-scale study, the National Human Activity Pattern Survey (NHAPS), was conducted by the U.S. EPA (Tsang and Klepeis, 1996). This phone survey was conducted over a 2-year period (1992–1994), during which 9386 households were surveyed in the 48 contiguous states. The participants were asked to recall all their

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activities during the preceding 24 h, and follow-up questions were asked to obtain additional information on chemical usage and other potential sources of chemical exposure. Specific information was obtained for both adults and children, including children too young to answer by obtaining the information from a resident adult. Although much demographic data and information were obtained, the questions of concern to this discussion centered around the time spent in various outdoor activities and where those activities occurred. The data indicate clear differences in activity patterns between various age groups, such as toddlers, children, and adolescents. There is a clear decrease in time spent outdoors for adolescents (13- to 17-year-olds) and adults in comparison to children (2–12 years). There are also differences in the activities, the intensity of the activities, and the duration and frequency of the activities among the various age groups. This means that if the goal were trying to simulate the activity patterns on turf in order to conduct exposure assessments, different routines would be required across several age groups. Alternatively, if the goal were trying to monitor people going about their normal routines, it would require monitoring several subgroups in order to ensure adequate data for each. Thus, assessing turf exposure is infinitely more complex than monitoring workers in an industrial or agricultural setting in which the routine is a fairly standardized task across all workers. Additional data have been generated to demonstrate the complexity of the activities that can occur on residential or recreational turf (Auyeung et al., 2006; Hore et al., 2006; U.S. EPA, 1997a; Xue et al., 2007). These data demonstrate that quantifying these activities is extremely difficult and time-consuming, and the data are probably obtained at great financial expense. It has been mentioned that dermal deposition is the primary route for the internal pesticide dosages in humans re-entering treated turf; inhalation is a minimal route by comparison. That discussion ignored the possible ingestion of pesticide residues that might occur from hand-to-mouth activities in very young children. Suffice it to say that this chapter is concerned with the dermal exposure to pesticides, and the hand-to-mouth exposures that might occur are beyond the scope of this discussion, even though they continue to be the subject of much current conjecture and research. The reader is referred to other sources for additional discussion of this topic (Ferguson et al., 2006; U.S. EPA, 1997a,b; Xue et al., 2007). Given the extreme complexity of attempting to monitor large numbers of people engaging in their normal activities on turf over the course of some extended period (or even for a single day), it is obvious that composing a routine that simulates these activities and condenses them into a fixed monitoring period is much easier in terms of cost, logistics, compliance, timing, etc. Of course, such considerations as routines for different age groups, simulating the frequency,

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intensity, and length of various age-related activities within a routine, and determining the overall length of the routines still represent an extremely difficult task. Such a simulated, or choreographed, routine would also have to be contactintensive to ensure that it was sufficiently conservative, or protective, for each age group. A method that has been used in indoor and outdoor exposure assessments to simulate contact with treated surfaces in a simple, reproducible, and temporally condensed program is the Jazzercise routine (Bernard et al., 2001; Formoli, 1996; Ross et al., 1990, 1991; U.S. EPA, 1997a,b). This routine ensures full body contact with the treated turf during an intense 20-min program. It is typically performed with subjects wearing whole-body dosimeters (i.e., long underwear), socks, and gloves (or it is performed bare-handed and hand wash samples plus face/neck wipes are collected from the exposed skin). This clothing scenario allows the calculation of potential dermal deposition of residue on an unclothed person, such as might be the case for a toddler. The exposure data are collected at the same time as TTR samples. From these data, a transfer coefficient is calculated (discussed later). ORETF (2009) took a unique approach to create an activity pattern routine to assess exposure potential to children. It used the NHAPS survey data for outdoor activities and analyzed them for nine age groups (e.g., 2 years, toddlers (2 or 3 years), 4–7 years, and 8–12 years) that were created based on physical and social development. The activities were divided into three intensity levels: passive (little or no contact with turf, such as playing on swings), active (nonaggressive contact with turf, such as playing with toys on turf), and hard direct contact (rigorous contact with turf, such as playing football or wrestling). The relative time of the three intensity levels, expressed as a relative ratio of the time engaged in each, was calculated for all age groups. Data collected in conjunction with the Stanford Research Institute, from which information was obtained on the frequency and duration of contacts of each body part of children with turf (Auyeung et al., 2006), were also used in constructing the final routine to ensure that it would incorporate enough contact of various parts of the body (e.g., hands, knees, and back) with the turf. From all these data, a composite 2-h routine was created for quantifying postapplication turf exposure to children. The routine incorporated more of the hard direct and less of the passive activities to ensure that it was sufficiently erring on the side of safety (i.e., tending to find higher exposure potential). The routine consisted of 12 activities or subroutines that were choreographed to ensure the proper intensity, duration, and frequency of turf contact with the various parts of the body. Regardless of the routine used, TTR samples must be simultaneously collected from the same plot of turf in order to relate exposure to source strength. The exposure data collected from passive dosimetry (micrograms of residue collected per unit time) and the TTR data (micrograms

of residue per unit area) are used to calculate a transfer co­efficient (TC; the contaminated area contacted per unit time). TC is the expression of a theoretical amount of contact that a person has with the pesticide-treated turf during a specific exposure activity. It is calculated as follows: TC (cm 2 /h) 

Dermal exposure (g/h) Transferable residue (g/cm 2 )

(1)

TC is a theoretical number because it assumes that 100% of the transferable residue is removed from only the amount of surface area that accounts for the micrograms of dermal exposure. This calculation is analogous to that for the renal clearance of a chemical.

46.4  Exposure assessment methodology Just as the amount of pesticide residue that transfers from the reservoir (i.e., turf) must be determined, the amount that is on the receptor (i.e., humans) must also be quantified. It is not the intent to provide a detailed discussion of the advantages and disadvantages of various methodologies here; we only mention the methods in common use to complete the discussion of exposure from treated turf. Biomonitoring data are well accepted by researchers and regulators for chemicals for which appropriate metabolism data are available (CDC, 2005, 2009; Paustenback and Galbraith, 2006; Sexton et al., 2004; U.S. EPA, 1997a, 1998). Biomonitoring data are often considered the gold standard of exposure data because they integrate the exposure across all possible routes despite the inherent differences in absorption rates across those routes. Noninvasive methods (e.g., urine collection) typically have higher compliance by participants and are less stressful for researcher and participant alike. However, passive dosimetry methods are even more widely used due to ease of implementation and ready acceptance by participants. Passive dosimetry methods used by many task forces involved in exposure assessment (e.g., Agricultural Re-Entry Task Force, ORETF, Non Dietary Exposure Task Force, Agricultural Handler Exposure Task Force, and Antimicrobial Exposure Assessment Task Force) involve whole-body dosimetry, wipes/washes of exposed skin surfaces such as face and hands, and use of air sampling pumps to assess potential inhalation exposure. These methods have the ability to determine the residues on various parts of the body that allow exposure estimates for various types of clothing and other activities (e.g., hand-to-mouth exposures for toddlers). The use of biomonitoring or passive dosimetry is accepted by the U.S. EPA (1998) and other regulatory agencies. Both methods are considered to provide roughly equivalent results given the typical variability found in these various types

Chapter | 46  Pesticide Exposure from Residential and Recreational Turf

of field studies (Bernard et al., 2001; U.S. EPA, 1997b). Although the passive dosimetry data generated by these methods are recognized as generalizable to other pesticides, the biomonitoring data are typically not.

46.5  Calculation of the safe   residue level Assuming that turf has been treated and has residues that may be transferred to humans resulting in some degree of potential exposure, how is the determination made regarding whether it is safe to re-enter the treated area? That is, when is a safe residue level (SRL) achieved? For treated turf, the timing of when an SRL must be achieved is very straightforward: It must be achieved just after the application. This is in contrast to the agricultural setting, in which the timing of the re-entry can be controlled by law (i.e., the label) to a time period such as 24 or 48 h after application. That is because re-entry in a residential setting cannot be controlled, and the assumption must be that re-entry could reasonably occur by children and adults very soon after the pesticide application. The toxicity of the active ingredient and its products is established by testing prior to entry into the marketplace. An appropriate toxicity endpoint and dose [typically a no-observed-effect level (NOEL) from the most sensitive study with an appropriate endpoint] are established together with a required safety factor (SF). Product-specific TTR data are generated, and the time zero residue level is calculated. However, the most difficult aspect of the risk assessment is to define the exposure parameter. Although toxicity of a product is constant, and its ability to transfer from turf over time is relatively constant, the exposure aspect can be highly variable, as described previously. Therefore, the use of a default TC attempts to simplify a very complex parameter. How all these inputs are used is described by the following equation:

2

SRL (g/cm ) 

NOEL (mg/kg)  1000 (g/mg)  body weight (kg) TC (cm 2 /h)  duration (h)  SF

(2)

If the SRL is higher than the calculated time zero residue level, then the product can be used in the manner in which it was tested (e.g., at the highest allowable application rate). If the SRL is lower than the time zero residue, then the product (i.e., active ingredient) use rate will have to be modified in some way to achieve the SRL (e.g., reduced application rate, reformulation, and watering in).

46.6  Regulatory approach to postapplication exposure to turf The U.S. EPA has developed guidelines for its approach to postapplication exposure assessment from treated turf in order

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to introduce a uniform means of calculating exposure (U.S. EPA, 1997a, 1999a,b). Following are some important default assumptions for the U.S. EPA’s tier 1 exposure assessment. The default TC used for exposure assessment on residential turf is actually derived from an indoor carpet study. It was obtained from a 20-min Jazzercise study done on indoor carpet (Formoli, 1996). The default TC is 46,000 cm2/h for adults and 8700 cm2/h for children. The TC is for an unclothed individual (i.e., to represent a child in a diaper). The Jazzercise routine is considered to be equal to 4 h of typical activity or 1 h of extreme activity. This is derived from comparison to studies in which scripted activities were performed by adults on turf and actual biomonitoring data were collected to determine the internal dose of the pesticide (Harris and Solomon, 1992; Vaccaro et al., 1996). The exposure for children (1–6 years old) is represented by toddlers (3 years old with an assumed weight of 15 kg). The surface area of an adult was scaled down to that for toddlers and a resultant TC of 8700 cm2/h was derived. The daily average time spent outdoors on residential turf for all age groups is assumed to be 2 h per day. This is derived from the 95th percentile for time on grass for both the 1- to 4-year-old and the 18- to 64-year-old groups (U.S. EPA, 1996). The transferable residue fraction from turf is assumed to be 5% of the application for the tier 1 assessment. Table 46.1 provides data from the U.S. EPA that contrasts their initial 20% transferable residue assumption with values actually obtained by various TTR methodologies. It is apparent why the initial 20% assumption was later decreased to 5%. Using these various assumptions, a daily exposure on the day of application would be calculated. Following is the U.S. EPA equation for performing this tier 1 exposure assessment:

PDR t - norm  (TTR t  CFI  TC  ET)/BW (3)

where PDRtnorm is the potential dose rate on day t (mg/day) normalized by person’s body weight; TTRt is the turf transferable residue on day t (g/cm2) [note that this was listed as DFR (i.e., dislodgeable foliar residue) in the U.S. EPA reference but changed here to TTR, which has been used throughout this chapter and is a more common usage term]; CF1 is the conversion factor to convert TTR units of micrograms to milligrams for the daily dose (0.001); TC is the transfer coefficient (cm2/h); ET is exposure time (h/day); and BW is body weight (kg). Using Eq. (3) and the following assumptions, a sample calculation is provided for the PDRtnorm. Following that calculation, the SRL (see Eq. 2) is then calculated for that product. The assumptions are as follows: a chemical with a shortterm dermal NOEL  10 mg/kg, maximum product application rate  1 lb ai/acre (equivalent to 11.2 g ai/cm2; 5% TR would be equivalent to 0.56 g ai/cm2) at the time of application, and a 100-fold uncertainty factor (UF) is required. Substituting into Eq. [3],

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PDR t - norm  (0.56 g/cm 2  0.001 mg/g  8700 cm 2 /h  2 h/day)/15 kg  0.65mg/kg/day



This exposure is higher than the acceptable daily exposure of 0.1 mg/kg (i.e., NOEL/UF). Using Eq. (2) again to calculate the SRL yields

2

SRL (g/cm )  

NOEL (mg/kg)  1000 (g/mg)  body weight (kg) TC (cm 2 /h)  duration (h)  SF 10 mg/kg  1000 g/mg  15 kg

8700 cm 2 /h  2 h  100  0.086 g/cm 2

(4)



Because the SRL is less than the TTR, this product does not pass its initial assessment and will require additional data such as that obtained from a product-specific TTR study to further refine the risk assessment. Other options, such as reformulation and adjustment of application rates, will be necessary if the product-specific TTR data do not demonstrate a SRL on the day of application. Recently, the U.S. EPA has used other assumptions to conduct these tier 1 assessments. For instance, a TC of 14,500 and 5200 cm2/h might be used for adults and children, respectively, assuming that shoes, shorts, and t-shirts are being worn while on the turf. Also, the actual transferable residue value from the TTR study can be used unless it is lower than 1% of that applied, in which case the default value of 5% transferable residue is used (U.S. EPA, 2001).

Conclusion It has been estimated that 67 million pounds of synthetic pesticides are used on lawns annually and that $700 million is spent on those pesticides (U.S. EPA, 2008). Approximately 20 million acres are cultivated as residential lawns, and lawn care is a $25 billion-a-year industry. When the combination of residential and recreational turf that is treated by homeowners and/or the lawn care industry is considered, it is obvious that the potential exists for human exposure from pesticide residues on turf. Developing methods to assess the transfer of residues from treated turf has been a priority of both regulatory bodies and industry. The modified California roller technique, based on that of Ross et al. (1990, 1991), has come to be accepted as the industry standard. Determining how to capture the potential contact of people with turf, given the myriad activities and contact intensity that vary across human developmental and age groups, is a monumental task. Multiple studies during approximately the past 20 years have used simulated play activities and the unique approach of a choreographed and highly reproducible

routine, Jazzercise. It is apparent that based on comparison of exposure data obtained by passive dosimetry and/or biomonitoring, Jazzercise provides a representative alternative routine to evaluate potential exposures from treated turf in future research. The U.S. EPA has produced standard operating procedures (SOPs) for use in performing exposure assessments for people re-entering treated turf. Since the initial use of these SOPs, much data have been generated that might result in modification of the original values. In October 2009, the U.S. EPA conducted another science advisory panel review of the residential SOPs to determine if they should be updated. This review will potentially affect future exposure assessments and future research in the field, although future research involving humans will also be tempered by the approach of the human studies review board.

References Auyeung, W., Canales, R. A., Beamer, P., Ferguson, A. C., and Leckie, J. O. (2006). Young children’s hand contact activities: an observational study via videotaping in primarily outdoor residential settings. J. Expo. Sci. Environ. Epidemiol. 16, 434–446. Bernard, C. E., Nuygen, H., Truong, D., and Krieger, R. I. (2001). Environmental residues and biomonitoring estimates of human insecticide exposure from treated residential turf. Arch. Environ. Contam. Toxicol. 41, 237–240. Black, K. G. (1993). “An assessment of children’s exposure to chlorpyrifos from contact with a treated lawn, doctoral dissertation,” Rutgers University, No. 9333377, UMI Dissertation Services, Ann Arbor, MI. Centers for Disease Control and Prevention (CDC) (2005). “National Report on Human Exposure to Environmental Chemicals,” Available at http:// www.cdc.gov/exposurereport (accessed May 2009). CDC, Atlanta. Centers for Disease Control and Prevention (CDC) (2009). “CDC’s Third National Report on Human Exposure to Environmental Chemicals,” Available at http://www.cdc.gov/exposurereport/pdf/factsheet_general. pdf (accessed May 2009). CDC, Atlanta. Cowell, J. E., Adams, S. A., Kunstman, J. L., and Mueth, M. G. (1993). “Comparison of foliar dissipation and turf dislodgeable residue sampling techniques. Pesticides in Urban Environments,” ACS Symposium Series 522, pp. 100–112. Driver, J. H., Ross, J. H., Pandian, M. D., Evans, J. B., and Whitmyre, G. K. (2001). Residential exposure assessment: An overview. In “Handbook of Pesticide Toxicology” (R. Krieger, ed.), Vol. 1, pp. 435–441. Academic Press, San Diego. Ferguson, A. C., Canales, R. A., Beamer, P., Auyeung, W., Key, M., Munninghoff, A., Lee, K. T., Robertson, A., and Leckie, J. O. (2006). Video methods in the quantification of children’s exposures. J. Expo. Sci. Environ. Epidemiol. 16, 287–298. Formoli, T. A. (1996). “Estimation of exposure of persons in California to pesticide products that contain propetamphos, HS-1731.” California Environmental Protection Agency, Sacramento. Fortune, C. R. (1997). “Evaluation of methods for collecting dislodgeable residues from turf, EPA/600/R-97/119.” U.S. EPA National Exposure Research Laboratory, Research Triangle Park, NC. Fuller, R., Klonne, D., Rosenheck, L., Eberhart, D., Worgan, J., and Ross, J. (2001). Modified California roller for measuring transferable residues on treated turfgrass. Bull. Environ. Contam. Toxicol. 67, 787–794.

Chapter | 46  Pesticide Exposure from Residential and Recreational Turf

Harris, S. A., and Solomon, K. R. (1992). Human exposure to 2,4-D following controlled activities on recently sprayed turf. J. Environ. Sci. Health B27, 9–22. Health and Environmental Sciences Institute (2004). “Residential Exposure Factors: Users Guide.” International Life Sciences Institute, Washington, DC. Hore, P., Zartarian, V., Xue, J., Ozkaynak, H., Wang, S. W., Yang, Y. C., Chu, P. L., Sheldon, L., Robson, M., Needham, L., Barr, D., Freeman, N., Georgopoulos, P., and Lioy, P. J. (2006). Children’s residential exposure to chlorpyrifos: Application of CPPAES field measurements of chlorpyrifos and TCPy within MENTOR/SHEDS-Pesticides model. Sci. Total Environ. 366, 525–537. Hurto, K. A., and Prinster, M. G. (1993). “Dissipation of turfgrass foliar dislodgeable residues of chlorpyrifos, DCPA, diazinon, isofenphos, and pendimethalin.” Pesticides in Urban Environments, ACS Symposium Series 522, pp. 86–99. Klonne, D., Cowell, J., Mueth, M., Eberhart, D., Rosenheck, L., Ross, J., and Worgan, J. (2001). Comparative study of five transferable turf residue methods. Bull. Environ. Contam. Toxicol. 67, 771–779. Lewis, R. G., Camann, D. E., Harding, H. J., and Agrawal, S. R. (1993). Comparison of transfer of surface chlorpyrifos residues from carpet by three dislodgeable residue methods. In “Measurement of Toxic and Related Air Pollutants.” Air & Waste Management Association, Pittsburgh, PA. Lewis, R. G., Fortmann, D. E., and Camann, D. E. (1994). Evaluation of methods for monitoring the potential exposure of small children to pesticides in the residential environment. Arch. Environ. Contam. Toxicol. 26, 37–46. Needham, L. L., Ozkaynak, H., Whyatt, R. M., Barr, D. B., Wang, R. Y., Naeher, L., Akland, G., Bahadori, T., Bradman, A., Fortmann, R., Liu, L. J., Morandi, M., O’Rourke, M. K., Thomas, K., Quackenboss, J., Ryan, P. B., and Zartarian, V. (2005). Exposure assessment in the National Children’s Study: Introduction. Environ. Health Perspect. 113, 1076–1082. Outdoor Residential Exposure Task Force. (2009). Personal communication from D. Johnson. Available at http://www.exposuretf.com/Home/ ORETF/tabid/58/Default.aspx (accessed April 2009). Paustenbach, D., and Galbraith, D. (2006). Biomonitoring: Is body burden relevant to public health? Regul. Toxicol. Pharmacol. 44, 249–261. Putnam, R. A., and Clark, J. M. (2007). Dosimetry and biomonitoring following golfer exposure to chlorpyrifos. In “Assessing Exposure and Reducing Risks to People from the Use of Pesticides” (R. I. Krieger, N. Ragsdale, and J. N. Seiber, eds.), ASC Symposium Series 951, pp. 157-171. ACS Books, Washington, DC. Rosenheck, L., Cowell, J., Mueth, M., Eberhart, D., Klonne, D., Norman, C., and Ross, J. (2001). Determination of a standardized sampling technique for pesticide transferable turf residues. Bull. Environ. Contam. Toxicol. 67, 780–786. Ross, J., Thongsinthusak, T., Fong, H. R., Margetich, S., and Kreiger, R. (1990). Measuring potential dermal transfer of surface pesticide residue generated from indoor fogger use: an interim report. Chemosphere 20, 349–360. Ross, J., Fong, H. R., Thongsinthusak, T., Margetich, S., and Kreiger, R. (1991). Measuring potential dermal transfer of surface pesticide residue generated from indoor fogger use: using the CDFA roller method. Chemosphere 22, 975–984. Sexton, K., Needham, L. L., and Pirkle, J. L. (2004). Human biomonitoring of environmental chemicals. Am. Sci. 92, 38–45. Thompson, D. G., Stephenson, G. R., and Sears, M. K. (1984). Persistence, distribution and dislodgeable residues of 2,4-D following its application to turfgrass. Pestic. Sci. 15, 353–360.

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Tsang, A., and Klepeis, N. (1996). “Descriptive Statistics Tables from a Detailed Analysis of the National Human Activity Pattern Survey (NHAPS) Data, EPA/600/R-96/074.” U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (EPA) (1994). “Pesticide Registration (PR) Notice 94-9: Announcing the Formation of Two Industry-Wide Task Forces: Agricultural Reentry Task Force and Outdoor Residential Exposure Task Force,” Available at http://www. epa.gov/PR_Notices/pr94-9.html (accessed April 2009). U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (EPA) (1995). “Data Call-In Notice.” U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (1996). “Exposure Factors Handbook, EPA/600/P-95/002Ba.” National Center for Environmental Assessment, Washington, DC. U.S. Environmental Protection Agency (EPA) (1997a). “Standard Operating Procedures (SOPs) for Residential Exposure Assessments,” Available at http://www.epa.gov/scipoly/sap/meetings/1997/september/sopindex. htm (accessed May 2009). U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (EPA) (1997b). “Meeting materials section (see passive dosimetry and biological monitoring documents),” Available at http://www.epa.gov/scipoly/SAP/meetings/2007/010907_ mtg.htm (accessed May 2009). U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (EPA) (1998). “Draft Series 875 – Occupational and Residential Exposure Test Guidelines, Group B: Postapplication Exposure Monitoring Test Guidelines, Version 5.4. of Prevention, Pesticides, and Toxic Substances.” U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (EPA) (1999a). “September 21–24, 1999: Issues Related to Standard Operating Procedures for Residential Exposure Assessment, LifeLineTM Project, Carbamate Pesticide, Issues Pertaining to Hazard and Dose Response Assessment and Review of American Cynamid Company’s Probabilistic Assessment for Chlorfenapyr,” Available at http://www.epa.gov/ scipoly/sap/meetings/1999/092199_mtg.htm#materials (accessed April 2009). U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (EPA) (1999b). “Overview of Issues Related to the Standard Operating Procedures for Residential Exposure Assessment,” Presented to the FIFRA Scientific Advisory Panel for the Meeting on September 21, 1999. Available at http:// www.epa.gov/scipoly/sap/meetings/1999/september/resid.pdf (accessed April 2009). U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (EPA) (2001). “Recommended Revisions to the Standard Operating Procedures (SOPs) for Residential Exposure Assessments,” Science Advisory Council for Exposure, Policy Number 12, Revised February 22, 2001. Office of Pesticide Programs, U.S. EPA, Washington, DC. U.S. Environmental Protection Agency. (2006). Federal Register: February 6, 2006, Vol. 71, No. 24, 40 CFR Parts 9 and 26, Federal Register Online via GPO Access [access.gpo.gov], DOCID:fr06fe06-7, Pg. 6137-6176. U.S. Environmental Protection Agency. (2008). “Landscaping with Native Plants.” November 26, 2008. Available at http://www.epa.gov/greenacres/wildones/handbk/wo8.html [citing data from Bormann, H., Balmori, D., and Geballe, T. (1993). Redesigning the American Lawn. Yale University Press, New Haven, CT] (accessed May 2009). Vaccaro, J. R., Nolan, R. J., Murphy, P. G., and Berbrich, D. B. (1996). The use of unique study design to estimate exposure of adults and children to surface and airborne chemicals. In Characterizing Sources of Indoor Air Pollution and Related Sink Effects (B. A. Tichenor, ed.), ASTM STP 1287, pp. 166–183. American Society for Testing and Materials, West Conshohocken, PA.

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Williams, R. L., Oliver, M. R., Ries, S. B., and Krieger, R. I. (2003). Transferable chlorpyrifos residue from turf grass and an empirical transfer coefficient for human exposure assessments. Bull. Environ. Contam. Toxicol. 70, 644–651. Williams, R. L., Bernard, C. E., Bigelow-Dyk, M., Ross, J. H., and Krieger, R. I. (2008). Measurement of transferable chemical residue

Hayes’ Handbook of Pesticide Toxicology

from nylon carpet using the California roller and a new megaCalifornia roller. J. Environ. Sci. Health B 43, 675–679. Xue, J., Zartarian, V., Moya, J., Freeman, N., Beamer, P., Black, K., Tulve, N., and Shalat, S. (2007). A meta-analysis of children’s handto-mouth frequency data for estimating nondietary ingestion exposure. Risk Anal. 27, 411–420.

Chapter 47

Lawn and Turf: Management and Environmental Issues of Turfgrass Pesticides J. Marshall Clark University of Massachusetts, Amherst, Massachusetts

Michael P. Kenna U.S. Golf Association, Far Hills, New Jersey

47.1  Turf environments 47.1.1  Usage and Functional Aspects 47.1.1.1  Usage The presence of turfgrass has become synonymous with the urban and suburban environments. As estimated by the National Home and Garden Pesticide Use Survey (Whitmore et al., 1993), approximately 83% of the estimated 84 million households in the United States are “urban” and 79% of these households have a private lawn. As of 2009, it is estimated that there are more than 16,000 golf courses in the United States that encompass an estimated 1.2 million acres of irrigated turfgrass, which is approximately 80% of the 1.5 million acres of maintained turfgrass that comprises rough, fairways, tees, and greens (Throssell et al., 2009). In addition, substantial turfgrass areas are associated with parks, athletic fields, gardens, cemeteries, public institutions, commercial properties, roadways, and sod farms. It is estimated that in the United States alone there are approximately 46.5 million acres of maintained turfgrass, an area larger than the total acreage for cotton, sorghum, barley, and oats (Joyce, 1998). Along with this increased use and maintenance of lawn and turf areas has come a heightened level of concern about potential human and wildlife health effects, drinking water pollution, and other environmental issues.

47.1.1.2  Functional Aspects A properly planned and maintained recreational turfgrass facility offers a diversity of functional benefits to the overall surrounding community, in addition to the physical and mental health benefits provided by the recreational activity. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

The following description of the functional benefits derived from turfgrass environments is condensed largely from Beard (2000). With additional research to develop best management strategies for turfgrass maintenance, the benefits derived from the turfgrass ecosystem and the environmental stewardship of these important recreational settings will coincide with each other to the mutual betterment of both. (a)  Soil and Wind Erosion Soil particles most susceptible to erosion are silt and clay. Clay particles are highly associated with phosphates, which contribute substantially to the eutrophication of ponds and lakes. Perennial turfgrasses are one of the most cost-efficient means to control water and wind erosion of soil. Dense, high-cut turfgrass reduces the overland flow of water such that runoff is insignificant in most instances so that turfgrass stands are not a significant source of soil particles entering bodies of water. Erosion control is due to the combined effects of a high shoot density and root mass for soil stabilization, plus a high canopy biomass, which resists lateral water flow and reduces water and wind velocity. (b)  Groundwater Recharge Turfgrasses conserve water by retaining potential water runoff, which allows more efficient vertical leaching. The leaf and stem biomass of turfgrass is substantial (1000– 30,000 kg ha1) and porous, which enhances surface water retention and subsequent infiltration. In addition, the turfgrass environment allows for a thriving earthworm population, which results in increased macropore structure and higher infiltration rates and water retention capacity. 1047

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(c)  Biodegradation of Organic Pollutants The turf environment supports an abundant and diverse group of soil microflora and microfauna, which are active and efficient decomposers of a variety of organic pollutants and pesticides. Thus, turf environments are characterized by an enhanced ability to capture surface water and associated pollutants and enhanced infiltration and bio­ remediation, all of which contribute to better groundwater protection. (d)  Soil Restoration Some of the world’s most fertile soils have developed beneath a turfgrass cover due to its high levels of live and residual biomass and subsequent efficient decomposition by its associated microflora and microfauna populations. In addition, regularly mowed turfgrass stands increase vegetative growth and the formation of dense, short, rapidly growing plants with a fibrous root system, rather than reproductive growth. This aspect, when coupled with the additional benefits mentioned previously, provides an efficient means for soil and land restoration as evidenced by the use of perennial turfgrasses in a variety of environmentally damaged areas, such as highly eroded rural land, mining operations, and garbage dumps. (e)  Heat Dissipation The high rate and efficiency of transpiration in turfgrass results in an increased ability to dissipate radiant heat and cool urban areas. In urban areas where low amounts of turfgrass are present, the overall ambient temperature can be as much a 5–7°C warmer than in adjacent suburban areas that are turfed. The transpirational cooling afforded by turfed environments results in lower energy needs for air conditioning adjacent homes and building, reduced electrical requirements, and less pollution due to the more efficient utilization of fossil fuels, etc. (f)  Wildlife Diversity In a properly designed “green” urban landscape, turfgrass areas, such as parks and golf courses, can coexit and even promote plant and animal diversity, natural habitats, and wetlands (Beard, 2000). In creating visually aesthetic turfgrass landscapes by the incorporation of lakes, pond, wetlands, trees, shrubs, flowers, etc., the designer concomitantly provides aquatic and terrestrial habitats that will attract a more diverse plant and animal wildlife population.

47.2  Turf pesticides and use 47.2.1  Turf Pesticide Industry The lawn and turf industry in the United States, including private and commercial lawn care, generates approximately

Hayes’ Handbook of Pesticide Toxicology

$25 billion annually (Joyce, 1998). The U.S. golf industry generates approximately $64 billion (Joyce, 1998). The expectations of home, institution, and business owners and golfers alike is high with respect to the quality of turf that they demand. Given the pest complex that is capable of damaging turf–disease, insects, nematodes, and weeds – a number of fungicide, insecticide, nematicides, and herbicide products, respectively, are applied to promote turf health. Currently, the application of pesticides is the major control factor in lawn and turf management, and this approach has been widely employed. The total urban consumer and professional pesticide market in 1991 was estimated at $2.2 billion (Racke, 2000). In 1996, the U.S. Environmental Protection Agency (EPA) estimated the urban consumer pesticide market at approximately $1.1 billion, indicating that the professional share of this market is approximately $1 billion (Joyce, 1998). According to Racke (2000), the pesticide market share in 1996 for turfgrass alone has been estimated at approximately $500–$700 million, including professional lawn care, golf courses, nurseries, institutions, landscapers, and sod farms. The U.S. EPA estimated in 1993 that 839 (76%), 193 (17%), and 73 (7%) million pounds of pesticide active ingredients were applied for agriculture, industrial/commercial/ government/golf course, and home use, respectively (Aspelin, 1994). This volume profile, when coupled with dollars spent, indicates that turfgrass pest control represents significant use of pesticides in the urban/suburban environment, including home lawns, parks, athletic fields, sod farms, commercial properties, and golf courses. For home and garden use, the U.S. EPA estimated that approximately 45 million pounds of herbicides, 12 million pounds of insecticides, and 4.5 million pounds of fungicides were applied in 1996 (Joyce, 1998). For golf courses, it is estimated that 55 pounds of pesticide active ingredients are applied to the average golf course each year (Joyce, 1998). The application of herbicides represents the largest pesticide use on an acreage basis, with between 1.4 and 3.2 million acres treated per year from 1993 to 1996 (Racke, 2000). Most herbicide application occurs on fairways and roughs. Fungicides are applied to golf courses, primarily to greens (73–93%), at higher frequencies and rates than other turf areas. Approximately 190,000 acres of golf course turf were treated with fungicides in 1996 (Racke, 2000). Insecticides are applied to tees, greens, and fairways, and little is applied to rough. In 1996, approximately 300,000 acres of golf course were treated with at least one insecticide application (Racke, 2000). On a comparison basis, as evidenced in a survey of pesticide use on Hawaiian golf courses, however, herbicides still predominate on an amount basis due to the overall large acreage treated and high application rates, resulting in approximately 80% of total pesticide use, followed by fungicides (16%), and insecticides (4%) (Racke, 2000).

Chapter | 47  Lawn and Turf: Management and Environmental Issues of Turfgrass Pesticides

Overall, treated yards and sod grass received more herbicide per acre but less insecticide and fungicide than most treated farms, including citrus fruits, other tree fruits, nuts, melons, vegetables, and cotton (Templeton et al., 1998). In terms of pounds of pesticide active ingredient per treated acre, golf courses have similar annual application rates as for citrus and tree fruits but exceed those for other treated farms and turfgrass usage.

47.2.2  Control of Turf Pests A substantial number of diseases, pests, and weeds damage turfgrass and impair its ability to function properly in the recreational activity that it was intended to provide. As mentioned previously, proper turfgrass management can minimize some of these problems in terms of the intensity and frequency of the damage. However, turfgrass is established and maintained usually with a single or few cultivars and in many cases in less than optimal environments.

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Environmentally stressed turfgrass is prone to disease and pest attack, and turf pest management is a necessity. Certainly for most golf courses and sports facilities, the application of pesticides is the accepted and required chemical control paradigm. Implementation of integrated turfgrass management practices and advances in alternative pest management strategies are providing new information on how best to manage turf facilities with minimal use of pesticides and reduced environmental hazard associated with their application. The following sections on disease, pest, and weed management are condensed largely from the Professional Guide for IPM in Turf (University of Massachusetts-Extension, 1999), 1998 Plant Protection and Fertilizer Usage Report [Golf Course Superintendents Association of America (GCSAA), 1998], and Racke (2000).

47.2.2.1  Diseases and Fungicides It is very difficult, if not impossible, to maintain turfgrass in highly managed settings without fungicides. Table 47.1

Table 47.1  Major Turfgrass Diseases, Pest Insects/Nematodes, and Weeds as Determined by Problem Intensity (10% Occurrence)a Rank

Disease problem and intensity

Insect/nematode problem and intensity

Weed problem and intensity

1

Dollar spot

39

Mole cricket

58

Annual bluegrass (Poa)

47

2

Pink snow mold

35

Fire ants

36

Goosegrass

38

3

Typhula blight

30

Turfgrass weevil

32

Crabgrass

36

4

Gray snow mold

27

White grub Japanese beetle

28

Clover, white

31

5

Fusarium patch

25

Cutworms

26

Dandelion

27

6

Anthracnose

24

Japanese beetle (adult)

24

Botton weed

13

7

Brown patch

24

Nematodes

20

Foxtail

12

8

Summer patch

23

Clover mite

18

Dichondra

11

9

Fairy ring

22

White grub Ataenius

17

Knotweed

10

10

Gray leaf spot

18

Crane fly

14

Mallow

10

11

Pythium blight

17

White grub European chafer

14

Plantain, broadleaf

10

12

Leaf spot

15

White grub June beetles

14

13

Mushroom and puffball

15

Sod webworm

14

14

Ophiobolus patch

14

Cicada killer

13

15

Pythium root rot

14

White grub masked chafer

13

16

Rhizoctonia blight

14

17

Helminthosporium

13

18

Take-all patch

13

19

Fusarium blight

10

20

Red thread

10

a

Data from Golf Course Superintendents Association of America (1998).

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lists the 20 most common turfgrass disease problems in terms of problem intensity. It is apparent that most turfgrass diseases are the result of fungi growth and thus can be effectively controlled using fungicides. Generally, fungicides are separated into two functional classes, the contact/ protectants and the penetrants, including the systemic types. Contact/protectant fungicides act at the plant surface and reduce infestation. It is necessary to reapply these types of fungicides every 5–14 days due to their environmental breakdown on the plant surface, removal due to wash-off (rain, dews, etc.), and removal by mowing. In addition, new growth, which is not covered, is susceptible to fungal attack. Spray tank adjuvants (spreader-stickers and surfactants) have been used to improve plant coverage and surface adherence, which are critical for effective protection. Table 47.2 lists some of the most widely utilized contact/protectant fungicides on turfgrass, their structures, and the diseases they commonly control. Penetrant fungicides are absorbed by the plant to provide their control. Locally absorbed fungicides, including chloroneb, iprodione, and vinclozolin, are absorbed into the leaf tissue and remain close to their point of entry. Other more broad-spectrum penetrant fungicides are absorbed by plant tissues and move away from the point of entry, generally in an upwards direction. Upwardly mobile penetrant fungicides include the benzanalides, benzimiazoles, sterol inhibitors (demethylation inhibiting), and strobilurins. Lastly, there are the more “systemic” penetrant fungicides for control of the oomycete diseases such Pythium blight. This group includes fosetyl-Al, metalaxyl, and propamocarb. Structures of each fungicide and the diseases that they control are given in Table 47.2. Penetrant fungicides usually maintain control for much longer periods of time (14–21 days or longer) compared with contact/protectant fungicides and can move into new tissues during growth of the plant. Of the major used fungicides on turf, the upwardly mobile penetrant types are the most chemically diverse (eight fungicides) and control the most fungal diseases (17 fungal diseases). Contact/protectant fungicides are the next most diverse group (four fungicides) and control 15 separate fungal diseases. Locally absorbed penetrant-type fungicides are represented by three fungicides, which control nine diseases, and the three systemic fungicides control two diseases, respectively.

47.2.2.2  Insects and Insecticides Table 47.1 lists the 15 most common insect pests on turf in terms of problem intensity in the United States (GCSAA, 1998). Pest insects include surface feeders that damage foliage by chewing, piercing, or sucking (e.g., armyworms, chinch bugs, cutworms, and sod webworms) and subsurface feeders, which attack roots and result in desiccation

Hayes’ Handbook of Pesticide Toxicology

and plants that are susceptible to drought (e.g., grubs). Mole crickets burrow through turf and can cause substantial disruption to the thatch/soil interface, which results in unsightly turfgrass surfaces. The type and intensity of insect damage is highly regional. Damage by Hyperodes weevils (annual bluegrass weevil) is commonly found on turf in the northeast United States, whereas fire ant and mole cricket damage is usually associated with the southeastern region. Insecticides such as acephate, bifenthin, carbaryl, chlorpyrifos, and spinosad are effective treatment for surface-feeding insect pests. Structures of each of these insecticides and the pests they control are given in Table 47.3. For subsurface insect pest control, insecticides such as acephate, chlorpyrifos, isazophos, and trichlorfon have provided good results. Halofenozide and imidacloprid, which both possess novel chemistry and no cross-resistance to previously used insecticides, appear to be highly effective and amenable to resistance and integrated pest management schemes. Acephate and isazophos have been effective against mole cricket damage, and fipronil, again with novel chemistry and no cross-resistance, provides an additional insecticidal tool for the management of this destructive pest, particularly on golf courses (see Table 47.3).

47.2.2.3  Nematodes and Nematicides There are two products commonly used as nematicides, fenamiphos (Nemacur) and ethoprop (Mocap), both of which are classified as organophosphates. Neither compound is registered for use on golf courses or turfgrass. Thus, the decision to apply these materials must be carefully evaluated and applications monitored judiciously. The presence of plant parasitic nematodes, by themselves, is not sufficient justification for application of nematicides. The threshold levels of nematode populations, which trigger the implementation of control applications, are dependent on various factors (e.g., time of year, distribution, and assay procedure used), show wide regional variations, and are usually based on limited experimental data. In addition, high populations of nematodes can coexist with limited damage on turf, making the justification of application difficult. If applied, however, nematicides should not be used where contamination of potable or groundwater may occur. Fenamiphos is a restricted-use material that is a systemic nematicide effective against ecto- and endoparasitic, free-living, cyst-forming, and root-knot nematodes and is recommended for application with and without soil incorporation. If used in zone II land areas (designated public wellhead), however, the turf manager should consult local pesticide regulations on this specific use. Ethoprop is a nonsystemic, nonfumigant nematicide, which is also effective against soil-dwelling insects. It can be used on Bermuda, St. Augustine, centipede, fescue,

Chapter | 47  Lawn and Turf: Management and Environmental Issues of Turfgrass Pesticides

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Table 47.2  Fungicides for Turfgrass Diseases Fungicide

Structure

Diseases controlled

Contact/protectant ChloroalkythioF

Brown patch, bipolaris, Drechslera leaf spot

O N

Captan (Orthocide)c

O

Dimethyldithiocarbamate Thiram (Defiant) Ethylenebisdithiocarbamate

SCCl3

S

CH3 H3C

N

S

N

S

S

CH3

CH3

[-SCSNHCH2CH2NHCSSMn-]Ar (Zn)

Mancozeb (Dithane) Tetrachloroiso-phtalonitrile Chlorothalonil (Daconil)

Brown patch, dollar spot, Fusarium patch, Typhula blight

Algae, anthracnose, brown patch, copper spot, dollar spot, Fusarium patch, Gray leaf spot, bipolaris, Drechslera leaf spot, red thread, rust, Typhula blight

CN Cl

Cl

Algae, brown patch, copper spot, dollar spot, downy mildew, Fusarium blight, Fusarium patch, necrotic ring spot, powdery mildew

CN

Cl Cl Penetrant Locally absorbed Chlorophenol

Pythium blight, Typhula blight

Chloroneb (Teremec)

Cl OCH3 CH3O Cl

Dicarboximides Iprodione (Chipco) Vinclozolin (Vorlan)

Brown patch, dollar spot, Fusarium patch, Fusarium blight, bipolaris, Drechslera leaf spot, necrotic ring spot, red thread, Typhula blight

O

Cl

N Cl

N

O

CH3 CH

O

Cl

N Cl

CONHCH(CH3)2

CH2

Brown patch, dollar spot, Fusarium patch, bipolaris, Drechslera leaf spot, red thread, Typhula blight

OCH(CH3)2

Brown patch, red thread, Typhula blight, yellow patch

O O

Upwardly mobile Benzanalide CONH Flutolamil (Prostar) Benzimidazole Thiophanate-methyl (Fungo)

CF3 NHCSNHCO2CH3 NHCSNHCO2CH3

Anthracnose, brown patch, copper spot, dollar spot, Fusarium blight, Fusarium patch, gray leaf spot, bipolaris, Drechslera leaf spot, necrotic ring spot, red thread, rust, smut, summer patch, take-all patch, yellow patch (Continued)

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Table 47.2  (Continued) Fungicide

Structure

Diseases controlled

Sterol inhibitors (DMIs) Cyproconazol (Sentinel)

CI

OH

CH3

C

C

CH2

H

N

Anthracnose, brown patch, copper spot, dollar spot, Fusarium patch, gray leaf spot, necrotic ring spot, powdery mildew, red thread, rust, smut, summer patch, Typhula blight

N

N Fenarimol (Rubigon)

OH CI

Anthracnose, brown patch, copper spot, dollar spot, Fusarium blight, Fusarium patch, necrotic ring spot, red thread, smut, summer patch, takeall patch, Typhula blight

N

C N CI

Myclobutanil (Eagle)

CN CI

C

(CH2)3CH3

CH2 N

Anthracnose, brown patch, copper spot, dollar spot, Fusarium patch, bipolaris, Drechslera leaf spot, necrotic ring spot, powdery mildew, red thread, rust, smut, summer patch

N

N Propriconazole (Banner)

CI

CI

O CH2

O

CH3CH2CH2

N

N

Anthracnose, brown patch, dollar spot, Fusarium patch, gray leaf spot, bipolaris, Drechslera leaf spot, necrotic ring spot, powdery mildew, red thread, rust, smut, summer patch, take-all patch, Typhula blight, yellow patch

N Triadimefon (Bayleton)

CI

O CH COC(CH3)3 N

N

Anthracnose, brown patch, copper spot, dollar spot, Fusarium blight, Fusarium patch, powdery mildew, red thread, rust, smut, summer patch, take-all patch, Typhula blight

N Strobilurin Azoxystrobin (Heritage)

Systemic Ethyl phosphate Fosetyl-Al (Alietle)

N O

N O CH3O

CN

(CH CH O 3

2

)

P O

3

Pythium blight Al

H

Phenylamide Metalaxyl (Subdine)

O

CO2CH3

Anthracnose, brown patch, Fusarium patch, gray leaf spot, bipolaris, Drechslera leaf spot, necrotic ring spot, Pythium blight, red thread, summer patch, take-all patch, Typhula blight, yellow patch

O CH3OCH2C CH3

N

CH3

Powdery mildew, Pythium blight

CHCO2CH3 CH3

Proplycarbamate Propamocarb (Banol)

(CH3)2N(CH2)3NHC02(CH2)2CH3 HC1

Pythium blight

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Table 47.3  Insecticides for Turfgrass Insect Pests Insecticide

Structure

Organophosphorous” Acephate (Orthene)c Chlorpyrifos (Dursban)rf

Insects controlled Fire ants, armyworms, cutworms, mole cricket, sod webworm

O CH3SPNHCOCH3 OCH3

Fire ants, cicada killer, army worms, billbug, chinch bug, cranefiy, cutworms, white grub masked chafer, white grub Ataenius, white grub European chafer, white grub Japanese beetle, adult Japanese beetle, sod webworm, turfgrass weevil

S Cl

N

OP(OCH2CH3)2

Cl Diazinon (Diazinon)

Cl CH3 N

OP(OCH2CH3)2 N

(CH3)2CH Fonophos (Crusade)

Chinch bug, white grub masked chafer, white grub Ataenius, white grub European chafer, white grub Japanese beetle, white grub June beetle

S

White grub masked chafer, white grub Ataenius, white grub European chafer, white grub Japanese beetle, white grub June beetle

S SPOCH2CH3 CH2CH3

Isazofos (Triumph)

(CH3)2CH

N

OP(OCH2CH3)2 N

Cl Isofenphos (Oftanol)

Cutworms, mole cricket

S

N

(CH3)2CHOCO

Cutworms, white grub masked chafer, white grub Ataenius, white grub European chafer, white grub Japanese beetle, white grub June beetle

S OP(OCH2CH3)2 NHCH(CH3)2

Trichlorfon (Pylox, Proxol)

Cutworms, white grub masked chafer, white grub Ataenius, white grub European chafer, white grub Japanese beetle, white grub June beetle

O ClCCHP(OCH3)2 OH

Carbamate Bendiocarb (Turcam)

O

CH3

O

CH3

White grub Ataenius, white grub Japanese beetle

CH3NH C O O Carbaryl (Sevin)

Pyrethroid lambda-Cyhalothrin (Scimitar)

Cicada killer, armyworms, cutworms, white grub Ataenius, white grub Japanese beetle, white grub June beetle, adult Japanese beetle, sod webworm

OCONHCH3

H

CI CF3

C

CH3

Cutworms, sod webworm

H CO2

CH CH3

C

CN

O

H

(Continued)

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Table 47.3  (Continued) Insecticide tau-Fluvalinate (Mavrik)

Structure

Insects controlled H

F3C

NH

C

CF3

Cyfluthrin (Tempo)

CH3

C

O

CO2CH

CH

Br C

CH3

CH

Br

F

Hyperodes weevil CO2 H

H

Hyperodes weevil, chinch bug, cutworm, sod webworm

CN

CH3

CH3

Deltamethrin (Deltagard)

Hyperodes weevil

CH3

H CO2CH2

CH

CI CI

CH3

H

CI

O

CH(CH3)2

CI Bifenthrin (Talstar)

Cutworms

CN

C CO2CH

C

CN

O

H

CH3

Amidinohydrazone Hydramethylnon (Amdro)

H N

CH3

N

CH3

Chloronicotinyl Imidacloprid (Merit)

N

CH C H

CF3

N N CH2 N

CI

N

N F3C

O S CF3

NH2

(CH3)2N CH3

Cutworms, mole cricket, sod webworms, turfgrass Atae-nius white grub masked chafer, white grub Ataenius, white grub European chafer, white grub Japanese beetle, white grub June beetle

CN

N Cl

O

O

O

CH3

O

OCH3 CH3

Army ants, cutworms, mole cricket, sod webworm OCH3 OCH3

HH

O O CH3CH2

Cutworms, sod webworms, turfgrass Ataenius white grub masked chafer, white grub Ataenius, white grub European chafer, white grub Japanese beetle, white grub June beetle

O

Halofenozide (Mach 2)

Cl

H

C(CH3)3

CONHN

Phenylpyrazole

Fire ants

Billbug, white grub masked chafer, white grub Ataenius, white grub European chafer, white grub Japanese beetle, white grub June beetle, turfgrass weevil

NO2

Diacylhydrazine

Spinosyn Spinosad (Conserve)

CF3

C

N H

CI

Fipronil (Chipcochoice)

H C CH

H H

O

H R

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Kentucky blue, perennial rye, and Bahia grasses in commercial turf. Currently, there is a soil fumigant (Curfew) that is registered for use on golf courses in Alabama, Florida, Georgia, Louisiana, Mississippi, North Carolina, South Carolina, and Texas. It has an active ingredient, 1,3-dichloropropene, which is highly volatile and must be applied by authorized operators.

herbicides need to be watered in and the soil should not be disturbed during the target weed seed germination interval. Control of emerged grass seedlings is usually achieved by the application of arsenates, such as MSMA, but these postemergence herbicides can injure turfgrass, particularly when applied during hot weather. Fenoxaprop (Acclaim) has been determined to be an effective replacement for the arsenates.

47.2.2.4  Weeds and Herbicides

47.2.3  Turf Pest Management

Turf can become infested with annual and perennial grasses (not the planted cultivar) and broadleaf plants that are controlled by the use of various herbicides. Table 47.1 lists the 11 most common turfgrass weed problems in terms of problem intensity. Control of nuisance aquatic weeds in lakes and ponds associated with parks and golf courses are also subjected to herbicide treatment. The control of grass weeds is primarily accomplished by the application of preemergence herbicides (e.g., benzamide, ethofumesate, pendimethalin, prodiamine, oxadiazon, bensulide, dithiopyr, and simazine). Broadleaf weeds are controlled predominantly using postemergence herbicides (e.g., Trimec mixtures of 2,4-D, dicamba, mecoprop, dicamba itself, and Confront mixtures of triclopyr plus clopyralid). The primary postemergence herbicides for grass weed control in turfgrass are the arsenates [e.g., methylarsonic acid (MSMA), disodium methylarsonate (DSMA), and amine methanearsonate (AMA)]. Control of annual and perennial weeds in nonplaying areas on golf courses and for turfgrass replacement is accomplished using nonselective postemergent herbicides such as glyphosate. These herbicides and some additional compounds used for weed control on turfgrass, along with their structures and effective weed targets, are given in Table 47.4. As of 1998, 2,4-D was still the most widely used herbicide on turf (26% of total active ingredient applied), followed by glyphosate (11%), mecoprop (9%), pendimethalin (8%), MSMA (8%), triazines (6%), chlorothalonil (5%), benfin (3%), dicamba (2%), and other herbicides (19%) (Aspelin and Grube, 1999). Preemergence herbicides are effective control agents for several weeks to months on most annual grass weeds. These materials have proven highly effective by providing excellent weed control with little or no injury to turf. For the most efficacious control of crabgrass and annual grasses, the preemergence herbicides need to be applied prior to the germination of the annual grass seeds. Thus, timing of application is critical and should be made 1–4 weeks before expected seed germination. Crabgrass germination occurs when the soil temperature is maintained at least at 65°F (18°C) for at least 1 week, and the monitoring of soil temperature is an efficient way to time the application. Following application, the preemergence

The control of disease, insect/nematode pests, and weeds on turfgrass is still largely accomplished through the use of pesticides. However, significant advances have been made, including adoption of integrated pest management (IPM) approaches and the introduction of new chemistry and biologically derived pesticides that are less environmentally stable and less toxic to humans and other nontarget organisms. IPM is based on the premise that less emphasis should be placed on the sole use of chemical control and that all available control means (chemical, biological, and cultural) be used in an integrated manner so that pesticides are used only when necessary and then only in a most efficient manner. The application of pesticides in IPM strategies is strategically timed and is based on monitoring or modeling and not on a calendar-based or prophylactic application schedule. Comprehensive summaries on the status and advances in turfgrass IPM are available (Leslie, 1994; Schumann et al., 1997). Although an impressive start has been made, overall adoption of IPM in the turfgrass management area has been slow. Of the practitioners in the major market segments (golf courses, home lawn care, and professional lawn care), golf course superintendents are the most active, with a survey indicating an 86% involvement in IPM practices and a 21% average reduction in pesticide use (GCSAA, 1998). However, these programs are not usually comprehensive, and major impediments to a wider adoption of turfgrass IPM approaches include a high expectation of golfers for nearly unblemished turf playing surfaces, high costs associated with existing pest monitoring protocols, and relatively small research and extension efforts (Potter, 1993). An important advancement for turfgrass IPM and resistance management schemes is the introduction of new and effective pesticide products via conventional means, which have novel chemistries, new modes of action, and high selectivity to insect pests. These include imidacloprid, introduced in 1995; halofenozide, introduced in 1997 for grub control; fipronil, introduced in 1996 for mole cricket control; and spinosad, introduced in 1997 for surface feeding insect control. Azoxystrobin, a broad-spectrum fungicide introduced in 1997, has been of great use in the professional turf disease market.

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Table 47.4  Herbicides for Turfgrass Weeds Herbicide

Structure

Weeds controlled

Cl

Annual bluegrass (Poa)

Preemergence   Benzamide” Pronamide (Kreb)c

CONHC(CH3)2CCH Cl

  Benzofuranyl alkanesulfonate

Annual bluegrass (Poa)

O OCH2CH3

Ethofumesate (Prograss)

CH3SO2O

  2,6-Dinitroanaline Pendimethalin (Weedgrass)

CH3

CH3

Annual bluegrass (Poa), crabgrass,

NO2 CH3

Prodiamine (Barricade)

NHCH(CH2CH3)2

CH3

NO2

H2N

NO2

F3C

foxtail, goosegrass

Annual bluegrass (Poa), crabgrass, foxtail, goosegrass

N(CH2CH2CH3)2 NO2

1,3,4-Oxidiazol Oxadiazon (Ronstar)

Crabgrass, goosegrass

O Cl O

Cl

N N

(CH3)3C

OCH(CH3)2

  Phosphoradithioate Bensulide (Betasan)   Pyridine Dithiopyr (Dimension)

Annual bluegrass (Poa)

S SO2NHCH2CH2SP(OCH(CH3)2)2)

F3C

CHF2

N

CH3SOC

Annual bluegrass (Poa), crabgrass, foxtail, goosegrass

COSCH3 CH2CH(CH3)2

  1,3,5-Triazine Simazine (Princep)

Cl N

Annual bluegrass (Poa)

NHCH2CH3

N N

NHCH2CH3

Postemergence   Benzofuranyl alkanesulfonate

Annual bluegrass (Poa)

O OCH2CH3

Ethofumesate (Progress)   Methylarsonic acid MSMA (Paconate 6)

CH3SO2O O CH3As(OH)2

CH3

CH3 Crabgrass, goosegrass

Chapter | 47  Lawn and Turf: Management and Environmental Issues of Turfgrass Pesticides

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Table 47.4  (Continued) Herbicide

Structure

 Phenoxyalkanoic and organic acids

Cl

Weeds controlled

Cl

2,4-D, dicamba, mecoprop (Trimec) Dicofop-methyl (Illoxam)

Botton weed, clover (white), dandelion, knotweed, dichondra, mallow, plantain (broadleaf)

OCH2CO2H

Goosegrass

CH3 O

Cl

O

CHCO2CH3

Cl Fenoxaprop (Acclaim)

Cl

CH3

O O N

Triclopyr, clopyralid (Confront)

O C

Crabgrass, goosegrass

CO2C2H5

H Clover (white), dandelion, dichondra, knotweed, plantain (broadleaf)

Cl N OCH2CO2H

Cl Cl   2,3,5-Triazine Simazine (Princep)

Cl

N N

NHCH2CH3

Dichondra

N NHCH2CH3

Nonselective postemergence   Glycerol phosphate Glyphosate (Roundup)

O

Annual perennial weed

HO2CCH2NHCH2P(OH)2

47.3  Turf pesticide issues 47.3.1  Environmental Issues A number of potential environmental issues have been raised due to the use of pesticides for turfgrass management. Of these, the impacts on surface and groundwater quality and the exposures to humans, wildlife, and other nontarget organisms are most critical. The following discussion summarizes research on the environmental fate, exposure, and best management practices for turf pesticides as reviewed by Balogh et al. (1992), Barbash and Resek (2000), Clark and Kenna (2000), Racke and Leslie (1993) and Thurman et al. (1992).

47.3.1.1  Significance to Drinking-Water Quality Most concerns for pesticide contamination of surface and groundwaters focus on their potential impact on drinkingwater quality. In rural areas, groundwater serves as the

major source for drinking water. In major metropolitan areas, however, as much as 95% of the drinking water comes from reservoirs. Pesticides have been determined to enter groundwater by leaching downward through the overlaying soil layers by chromatographic movement through soil micropores and/or macropores. Pesticide contamination of surface water has been shown to occur by spray drift following application, by rain and dust-out events, and by runoff water processes, including soluble pesticides and bound residues on soil particulate during erosion. As summarized by Barbash and Resek (2000), approximately 90 different pesticides and transformation products have been detected in groundwater in the United States. With few exceptions, pesticide concentrations exceeded the maximum contamination levels (MCLs) or lifetime health advisory levels (HALs) in less than 1% of the sources sampled. Pesticides that exceeded the MCL or HAL crit­ eria were from areas of known or suspected contamination where extensive sampling had taken place. These include

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aldicarb and its sulfoxide and sulfone transformation products, arsenic, 1,2-dibromo-3-chloropropane (DBCP), and 1,2-dibromoethane (EDB). The preceding summary indicates that the concentrations of most agricultural pesticides (~99%) are below existing drinking-water criteria established for the protection of human health. There are a number of caveats, however, that must be considered prior to passing judgment on the extent and significance of pesticide contamination of drinking water. First, MCLs and other water-quality indicators do not exist for all pesticides and most transformation products. Second, MCLs and other criteria do not take into account the possible cumulative effects that may occur in mixtures, which is likely the way most exposures take place. Third, analysis for many pesticides and most transformation products is incomplete with limited sampling in urbanized areas, where pesticide use is often high. Indeed, nonagricultural pesticide use on lawns and golf courses in the United States is substantial (estimated at approximately 25 million kg annually) and comparable to the amounts used in many agricultural settings. Thus, it is reasonable to assume that turfgrass pesticides may have impacted the quality of drinking water to a similar extent as determined for pesticides used in agricultural settings. In support of this contention, dacthal (DCPA – a herbicide used primarily in nonagricultural settings) was detected most frequently in groundwater during the U.S. EPA’s national pesticide survey (1990). A more complete evaluation of the impact of turf pesticides on drinking-water quality has begun to emerge. (a)  Surface Water The potential for contamination of surface water by turf pesticides is of concern primarily to wildlife and other nontarget organisms, particularly humans. Research on the potential of surface water contamination by pesticides used on turf via runoff processes has indicated that limited transport is usually detected under most conditions (Racke, 2000). The amount of pesticide that is transported from the turfgrass environment, however, is much reduced compared with that which occurs on bare soil or in cropping situations. Gold et al. (1988) first reported a low level of pesticides associated with runoff from sloped turfgrass plots. Concentrations of the herbicides 2,4-D and dicamba ranged from nondetectable to 15 and 38 ppb, respectively. The distribution of concentrations was skewed to the lower concentration values, and the large majority of the samples were below 1 ppb. The levels of 2,4-D and dicamba detected in the percolate at the highest level of treatment were 0.4 and 1.0%, respectively, of the total applied. In a similar study that evaluated a wider range of turf pesticides (chlorpyrifos, dicamba, 2,4-D ester, 2,4-DP ester, and pendimethalin), Harrison et al. (1993) found that following irrigation of 150 mm h1 for 60 min, which produced

Hayes’ Handbook of Pesticide Toxicology

average runoff values of 0.8, 13.4, and 11.6% of the total water applied, no residues of pendimethalin, chlorpyrifos, or the esters of 2,4-D and 2,4-DP were detected in any sample. Mean concentrations of 2,4-D acid and 2,4-DP acid and dicamba for individual events ranged as high as 312, 210, and 252 ppb, respectively. Nondetections for the same pesticides accounted for 63, 64, and 51% of the samples, and another 30, 25, and 47% of these samples were less that 70 ppb, the MCL for 2,4-D. Most interestingly, no detectable levels of the highly sorbed pesticides chlorpyrifos and pendimethalin and only low levels of the more water-soluble pesticides 2,4-D and dicamba (0.8 and 1.6% of total applied) were found in runoff following a 60-min irrigation event that mimicked a 100-year-frequency storm. Shuman et al. (2000) evaluated the potential movement of pesticides in runoff water following application to Bermuda grass plots that were sloped at a 5% angle. The fraction of water leaving the plots at 24, 48, 96, and 192 h following treatment was 45, 72, 40, and 36%, respectively. Only samples collected during the first 192 h following treatment had pesticide concentrations above the detection level (1 ppb). The highest concentrations of pesticide in runoff water were detected during the first rainfall effect, which was applied 24 h following treatment. 2,4-D, dicamba, and mecoprop treatments resulted in the highest concentrations in runoff water and were 800, 360, and 810 ppb, respectively, accounting for 9.6, 14.6, and 14.4% of the respective total pesticides applied. The levels of chlorothal­ onil (290 ppb), dithiopyr (39 ppb), chlorpyrifos (19 ppb), benefin (3 ppb), and pendimethalin (9 ppb) were substantially reduced and accounted for 0.8, 1.9, 0.1, 0.01, and 0.01% of the respective total pesticide applied. These authors concluded that the relationship between the fraction transported and the negative log of the analyte solubility in water was best described using a quadratic rather than a linear equation, possibly due to the precipitation and crystallization of some analytes resulting in reduced dissolution of analyte molecules. In studies that comparatively evaluated the runoff of a fungicide (triadimefon), a herbicide (mecoprop), and an insecticide (isazofos) from sloped turfgrass plots, Watschke et al. (2000) found similar results in that these pesticides, when applied properly to established perennial ryegrass and creeping bentgrass managed as fairway-type turf, were detected at only low levels even though runoff had been forced by unrealistically high irrigation events (152 mm h1). Most interestingly, however, was the substantial reduction in pesticide residues in runoff water as the turfgrass plots matured. In the initial year of the study, the experimental plots were newly established and more representative of a grow-in situation. Residues of mecoprop and triadimefon, which were collected in the first liter of runoff water at 24 h following application in each of the 2 years of the study, were reduced 85% or more in the second year of the study. This finding clearly indicates the vulnerability

Chapter | 47  Lawn and Turf: Management and Environmental Issues of Turfgrass Pesticides

of newly established turf for pesticide contamination of surface waters due to runoff events and the need for judicial use of irrigation protocols during this period. In conclusion, pesticide contamination of surface water following application to turfgrass can occur in most circumstances, but residue levels are low and do not usually exceed existing action levels for drinking water in the United States. The overall result of a surface water monitoring study involving a number of North American golf courses was that widespread and/or repeated water quality impacts by pesticides due to golf course management practices were not currently occurring (Cohen et al., 1999). The average concentration of pesticides in surface water ranged from 0.07 to 6.8 ppb. The 95th percentile concentrations were between 0.015 and 10 ppb. Eleven of the 31 pesticides detected in surface waters exceeded the MCL/HAL values and/or the maximum allowable concentration values (MACs) for aquatic organisms and included acephate, chlorothalonil, chlorpyrifos, diazinon, ethoprop, fenamiphos sulfoxide and sulfone, isofenphos, malathion, methamidophos, and simazine. Although water quality standards were exceeded in a few sites (individual pesticide database entries that exceeded MCL or HAL values for surface water were 0.29% of the total), no toxicologically significant impacts were noted. Similar findings have been determined for Japanese golf courses (Morioka and Cho, 1992; Tomimori et al., 1994). The low level of pesticide transport due to runoff events in turfgrass appears to due to the unique features found in the turf environment and includes a low inherent level of soil erosion due to reduce water flow and dense stem and root structures; increased sorption of organic compounds, including many pesticides, by the thatch layer; and increased infiltration rates associated with soils that have an overlaying turfgrass planting. Nevertheless, heavy textured, compacted soils and moist soils are much more prone to runoff losses than drier soils, and the application of water-soluble herbicides on dormant turf can result in high levels of runoff loss. (b)  Groundwater Contamination of groundwater by pesticides used to manage turfgrass is a major concern from a human drinkingwater quality standpoint. A well-managed turf is an efficient system for the retention and degradation of ­pesticides. Its edaphic environment, described ­ previously, greatly enhances the biota and biodegradation of organic pollutants, including many pesticides. The high surface area and organic content of the thatch/mat layers also provide a major reservoir for the retention of organic compounds, which reduces leaching and macropore transport rates. For many pesticides, their application to turf results in increased rates of degradation and a reduction in leaching compared with bare soils. The half-lives for pendimethalin, isazofos, chlorpyrifos, and metalaxyl in bare soil are 90, 34, 30, and 70 days, respectively. Under turf

1059

conditions, they become 6–15, 5–11, 6–12 and 10–25 days, respectively (Horst et al., 1996). An example of the extent of this attenuation is given by isazofos. When applied to bare soil, the ground ubiquity score (GUS) for isazofos has been determined to be 3.06 with a field halflife of 34 days (Wauchope et al., 1992). Using the half-life value determined by Horst et al. (1996), the GUS value of isazofos applied to turf is calculated to be between 2.08 to 1.40, with a half-life between 11 to 5 days (Branham et al., 2000). GUS values between 1.8 and 2.8 indicate that a pesticide may be a leacher and its use could lead to ground­ water contamination. GUS values of 1.8 or less indicate a low likelihood of groundwater contamination by leaching. In the case of isazofos application to turf, there is moderate to low concern that it will appear in groundwater. However, research to date has established the general finding that limited transport of pesticides applied to turf does occur in most situations (Racke, 2000). Pesticide leaching by macropore flow most likely represents the pathway for most pesticides applied to turf to reach groundwater (Branham et al., 2000). A particularly vulnerable site for pesticide leaching and potential groundwater contamination is the golf course green. Many greens are constructed for maximum infiltration and percolation of water via the rooting mix. Root zone mixtures are composed of at least 85% sand by volume, which allows rapid water percolation and a very low cation exchange capacity. Soil sterilization is also recommended during construction to reduce weed and disease problems. This practice, however, reduces the microbial populations and the usually high level of pesticide biodegradation in these areas of the golf course. Together, these characteristics can facilitate the rapid movement of un­altered pesticides through the rooting mix and into effluent water from the green, which can ultimately be a source of contamination to groundwater. The mobility and persistence of pesticides in the golf course green is influenced by the amount and composition of soil used for their construction. The United States Golf Association (USGA) has provided recommendations that specify the construction of largely sand-based greens to ensure a predicable putting surface. The USGA green drains rapidly but still provides an acceptable level of water retention and root zone depth. Initially, USGA greens were to have at least 20% organic matter, which should help in retarding the movement of pesticides to groundwater. In hot climates, however, considerable percolation can occur on greens, particularly in the presence of extensive irrigation. Nevertheless, turfgrass systems are relatively well protected in terms of groundwater contamination due to the leaching of pesticides. The final distribution of a pesticide in the turf environment is highly dependent on the type and application rate of the pesticide used, the method of application, and the rate and timing of postapplication irrigation. Because of these variables, subsurface movement and

1060

groundwater contamination by pesticides used in turf management have been determined despite the positive effects that turf environments display. Cisar and Snyder (1993, 1996) investigated the leaching of organophosphorous pesticides via a root zone constructed of sand and percolate samples collected at a depth of 20–26 cm. A maximum concentration of fenamiphos (8 ppb) was detected in the percolate following an initial application rate of 11.25 kg ha1. The two oxidative metabolites of fenamiphos (fenamiphos sulfoxide and sulfone), however, reached concentrations as high as 2300 ppb. A total of 17.7% of the applied fenamiphos was detected in the percolate as the sulfoxide and sulfone metabol­ites following the initial application. Following a second application, only 1.1% recovery of these metabolites in the percolate was found, indicating a rapid and extensive enhanced metabolism of fenamiphos and its oxidative metabolites in this constructed root zone mixture. The fate of DCPA and isazofos has also been studied using model turf ecosystems, which consisted of a glass atmospheric chamber, a brass base to hold the turf sample with a porous ceramic tension plate in the bottom to simulate normal soil water movement, and chemical traps to collect 14C-labeled compound leaving the system (Branham et al., 1993). Of the six soil type/irrigation frequency treatments tested, only the sand soil with turf/thatch that was irrigated every fourth day had less than 50% of the applied DCPA remaining after 8 weeks. All other treatments had between 67 and 95% of the applied DCPA remaining. Between 1 and 25% of the applied DCPA was detected as DCPA acid metabolite in the leachate. In similar experiments, the presence of turf/thatch layers resulted in enhanced degradation of isazofos with only 5% of the total applied remaining, whereas 13% remained in bare soil. The major metabolite of isazofos, which is formed by hydrolytic ester cleavage, accumulated rapidly in all soil types and accounted for 20–60% of the total isazofos applied 1 week after application. The hydrolytic metabolite was more mobile than isazofos and accounted for 8–17% of the total isazofos applied in the leachate compared with 1–5% isazofos at 4 weeks postapplication. These initial studies clearly demonstrate the ability of pesticides to leach following their application to turf, their enhanced degradation in the presence of turf, and their increased retention in the presence of thatch. However, such model systems rarely mimic the natural environment. In particular, both systems had very shallow soil profiles compared with that found in natural turf settings. To this end, recent studies have monitored turf pesticides in more undisturbed or natural settings. Using intact-core monolith lysimeters, Branham et al. (2000) reported that of the seven pesticides applied (2,4-D, chlorothalonil, fenarimol, isazofos, propiconazole, metalaxyl, and triadimefon), only triadimefon was detected in the percolate during the course of the 2-year study.

Hayes’ Handbook of Pesticide Toxicology

Triadimefon was detected multiple times at both lysimeters, and its concentration in the percolate ranged from just above the level of detection to as high as 32 ppb. The leaching of triadimefon was considered to be by macropore flow and correlated positively with major rainfall events. The conversion of triadimefon to its more mobile metabolite, triadimenol, was not determined. However, Petrovic et al. (1993) reported the leaching of triadimefon and a greater amount of triadimenol in the leachate. On native Kentucky bluegrass fairways, Schumann et al. (2000) found a similar rapid and extensive conversion of triadimefon to triadimenol, but both were retained in the thatch layer almost completely during the 28-day experimental interval. The potential for pesticide leaching through golf course putting green surfaces also has been studied. Initially, greenhouse lysimeters were developed to study the fate and movement of turf herbicides applied to Bermuda grass that was maintained as a green (Smith and Tillotson, 1993). Following the application of the dimethylamine salt of 2,4-D, only trace amounts of the herbicide were detected in the leachate (2 ppb). Shuman et al. (2000) investigated the leaching of pesticides applied to simulated golf course greens where benefin, chlorothalonil, chlorpyrifos, dicamba, dithiopyr, 2,4-D, mecoprop, and pendimethalin were applied to Bermuda grass and bentgrass plots with a 5% slope. Only low concentrations of these pesticides were detected in the leachate from lysimeters beneath these simulated greens. When present, residues were detected only during the 4- to 6-week interval following application. Only dicamba (1.2–2.6 ppb), dithiopyr (2.4–1.6 ppb), chlorothalonil (10.5–15 ppb), and its hydroxylated metabolite, HO-chlorothalonil (39–43 ppb) were detected. Less than 0.5% of any applied pesticide was present in the lysimeter leachate during the 70-day experiment. From 1991 to 1994, the fate and persistence of fenamiphos, fonofos, chlorpyrifos, isazofos, isofenphos, ethoprop, 2,4-D, and dicamba were also investigated following application to a USGA green (Cisar and Snyder, 2000). Although all of the organophosphorous pesticides were detected in leachate, most were below a concentration of 0.7 ppb. The leachate concentration of fenamiphos ranged from 1.56 to 2.06 ppb and ethoprop was 1.46 ppb. The combined concentration of the sulfoxide and sulfone metabolites of fenamiphos ranged from 46 to 585 ppb. The concentration for 2,4-D varied from 1.2 to 2.6 ppb and from 1.7 to 2.5 ppb for dicamba, respectively. In terms of the percentage of total applied, the fenamiphos metabolites made up 1–18%, dicamba 9.7–10.8%, and 2,4-D 0.5–1.6% of the total applied. The other pesticides were less than 1.0% of their respective total. The leaching data are consistent with concurrently collected data for thatch and soil. Because in most cases pesticides were mostly found in thatch and little move into the underlying soil, it was expected and determined that only small quantities were detected in percolate

Chapter | 47  Lawn and Turf: Management and Environmental Issues of Turfgrass Pesticides

waters. When less pesticide was detected in thatch, correspondingly more was found in the percolate. Current information on the mobility, fate, and level of groundwater contamination by turf pesticides is limited, and monitoring is variable and not widespread. Nevertheless, the immerging picture indicates that although detections have been reported, the frequency of detection and the levels of residues are generally lower than those for surface waters, and the frequency of exceedence of drinking-water criteria (e.g., MAC, MCL, and HAL) is extremely rare (Cohen et al., 1999). Cohen et al. (1990) first reported pesticide contamination of 16 monitoring wells on golf courses and found chlorpyrifos, 2,4-D, dicamba, isofenphos, and trichloropyridinol (TCP) in one of the wells, chlorothalonil in two wells, DCPA in three wells, heptachlor epoxide in four wells, chlordane in seven wells, and DCBA in nine wells. Of the wells with detectable pesticide residue levels, 80% had concentrations below 5 ppb. The results from this research demonstrated a positive relationship between the rates of pesticide application and the frequency of pesticide detection in groundwater below different areas on the golf course, such as tees, greens, fairways, and untreated locations. Prometon, a nonselective, nonagricultural herbicide, was found in substantially higher frequency (80%) than other herbicides (36%) in wells located within 400 m of private residences or within 3.2 km of golf courses during the Midcontinent Pesticide Study (MCPS) (Burkart and Kolpin, 1993, Kolpin et al., 1994). In addition to these studies, prometon has also been detected in ground­water in many different regions of the United States, including drinking-water wells monitored during the National Pesticide Survey (U.S. EPA, 1990) and in groundwater in Florida, California, New Jersey, Illinois, Nebraska, Oklahoma, and Nevada (Barbash and Resek, 2000). The most commonly detected pesticide-related compounds were from the hydrolytic degradation of DCPA (dacthal), a widely used herbicide on golf courses, as well as in urban areas and in agriculture. TPA, produced during the hydrolysis of DCPA, was the most frequently detected compound identified during the 1992 MCPS. These observations are consistent with the detection of DCPA and its hydrolytic degradation products in groundwater both by the review process and in matrix-distribution studies that examined their behavior following application to turf environments. Interest in groundwater contamination by pesticides and fertilizers applied to golf courses and other turf environments has increased significantly since the early 1990s due mostly to local permitting processes. An extensive evaluation of the water quality impacts by chemicals used to manage golf courses in the United States was undertaken by Cohen et al. (1999). In this study, which evaluated the groundwater quality of 36 individual golf courses (Table 47.5), 21 pesticides and metabolites were identified in 160 samples that had detectable levels of pesticide residues

1061

(1.3% of the groundwater organic chemical database entries). Of these, only nine detections (0.07% of the total organic chemical database entries) exceeded the MCL or HAL drinking-water health advisory criteria, and most were associated with golf courses in Florida. The pesticides responsible for the nine exceeded values were arsenic (4), chlordane (2), acephate (1), atrazine (1), and bentazon (1). The average concentration of pesticides in ground­ water was 0.09–3.6 ppb, and the 95th percentile concentration was between 0 and 10 ppb, depending on the level of nondetection used. A survey evaluated the presence of pesticides used in urban landscape maintenance in residential drinking-water wells (Eitzer and Chevalier, 1999). The survey included the analysis of 19 pesticides from water samples taken from 53 residential drinking-water wells in a single town in Connecticut. Six of the 53 wells had detectable levels of at least one of the pesticides examined. The frequency of pesticide occurrence (11%) is less than that reported for surface aquifers in the Connecticut, Housatonic, and Thames river basins (80% in agricultural, 60% in urban, and 48% in undeveloped areas) (Kolpin et al., 1998). Pesticides detected were diazinon, DCPA, trifluralin, lindane, chlorpyrifos, and chlordane, and their concentrations ranged from 0.02 to 0.22 ppb. Thus, none of the pesticides detected in these residential drinking-water samples exceeded any MCL for drinking water or reference dose (RFD) established by the U.S. EPA, assuming a 50-kg individual and a 2-liter water consumption rate. Thus, specific instances in which pesticides are detected in groundwater underlying turf environments at concentrations of concern are limited and usually associated with unusual conditions or management practices. For instance, the detection of chlordane and heptachlor, two cyclodiene insecticides of particularly low water solubility, in groundwater beneath a limited number of Cape Cod golf courses at levels that exceeded the HAL values has been attributed to repeated, heavy applications and a preferential flow of bound particulate phase (Cohen et al., 1990). In addition, the presence of trace levels of DCPA reported in a U.S. EPA study of well samples was correlated to its extensive use on golf courses and commercially maintained landscaping in urban areas (Racke, 2000). In summary, pesticide leaching studies have determined that dense turf cover reduces the extent for leaching of pesticides and the potential for groundwater contamination. Conversely, more leaching will occur in newly established or poorly maintained turf stands. Generally, sandy soils are more vulnerable to leaching losses than are more clayey soils. The physical and chemical properties of pesticides are good indicators of their leaching potential. Lastly, current models for the prediction of pesticide loss due to leaching and the impact that this aspect has on ground­ water quality usually overpredict the extent of the problem, particularly if valid adjustments are not introduced to

Hayes’ Handbook of Pesticide Toxicology

1062

Table 47.5  Pesticides, Metabolites, and Solvents Analyzed in Groundwater Monitoring Studies on Golf Courses Analytes analyzed in one or more samples Analyte

GW

1,1-Dichloroethane 1,1-Dichloroethylene

SW

Analyte

GW

SW

Analyte

GW

SW



DDD





Malathion







DDE





Mancozeb



1,1,1-Trichloroethane



DDT





Merphos



1,1,1,1-Tetrachloroethane



Daethal

1,1,2,2-Trichloroethane



Daethal diacid

1,2-Dichloroethylene



1,2-Dichlorobromoprop 1,2-Dichlorobenzene



Metalaxyl







Methamidophos





Dalapon



Methiocarb





Delta-BHC





Methomyl





Demeton-O



Methoxychlor



1,2-Dichloroethane



Demeton-S



Methyl bromide



1,2-Dichloropropane



Diazinon





Methyl isothiocyanate



1,3-Dichlorobenzene



Dicamba





Methyl parathion



1,3-Dichloropropene



Dichlorvos



Methyl chloride



1,4-Dichlorobenzene



Dichlorprop





Metribuzin





2,4,5-T



Dieldrin





Mevinphos





2,4,5-TP



Disulfoton





Naled





2,4-Da



Diuron





Norflurazon









2,4-DB





Endosulfan I





Oryzalin





Acephate





Endosulfan II





Oxamyl





Acetic acid



Endosulfan sulfate





PCNB







Endrin





Phorate







Endrin aldehyde





Picloram







Prodiamine

Aldrin Alpha-BHC





Endrin ketone







Ethion



Alpha-chlordane Ametryn



Prometryn





Pronamide





Propanil

Anilazine





Ethofumesate

Arsenic





Ethoprop



Atrazine





Ethyl parathion



Propiconazole





Ethylbenzene



Propoxur





Ethylene dibromide



Ronnel

Fenamiphos



Azinphos-methyl Bendiocarb



Benomyl



Bentazon



Benzene



Beta-BHC Bromacil



   

Siduron

 



Fenamiphos sulfone





Simazine





Sulprofos





Fenamiphos sulfoxide Fenarimol





Tetrachlorvinphos





Fensulfothion



TCP



Carbaryl





Fenthion



Terbufos



Carbofuran





Gamma-chlordane

Carbon tetrachloride



Chlordane



Chlorobenzene

   



Tetrachloroethylene



Glyphosate





Thiram



Heptachlor





Tokuthion



Heptachlor epoxide





Toluene



Chloroethane



Hexazinone





Toxaphene





Chloroform



Iprodione





Triadimefon





Chloropicrin



Isofenphos





Trichloroethylene



Chlorothalonil





Lindane





Trichloronate

Chloropyrifos





Linuron





Triclopyr

Chloropyrifos ethyl





MCPA





Xylenes



MCPP





Coumaphos a



 Underlining indicates analytes detected in GW and/or SW in one or more samples. GW, groundwater; SW, surface water. Adapted from Cohen et al. (1999), Table 5.3, p. 803.

 

   



Chapter | 47  Lawn and Turf: Management and Environmental Issues of Turfgrass Pesticides

account for the role of the turf canopy and thatch/mat layer (Kenna and Snow, 2000).

47.3.2  Exposure Issues 47.3.2.1  Humans There has been and will continue to be great concern over human exposure considerations related to the use of pesticides in the management of turf environments. Given the intensity of pesticide use, the extent of activities and time spent on turf, and the exposure potential for infants, children, and adults alike, this is expected and germane. A large amount of time and effort has been expended in the determination of applicator exposure issues and means to mitigate problematic exposure situations. However, there are potential exposure concerns for all who reenter turfgrass areas following pesticide applications. The primary route of potential reentry exposure involves dermal uptake from dislodgeable foliar pesticide residues (pesticide residues available by contact or abrasion for skin absorption) present on treated turfgrass foliage. Because turfgrass has a very dense canopy, it is expected that a larger proportion of the applied pesticide will remain on the turfgrass leaves compared with an agricultural cropping situation in which a substantial proportion of the pesticide is applied directly to the soil surface. Thus, dermal exposure to dislodgeable residue in the turfgrass environment is expected to be significant. However, most turfgrass cultivars used for lawns, golf courses, etc. have substantial waxy layers associated with their blades and all grasses produce organically rich thatch/mat layers, both of which act as efficient reservoirs for lipophilic pesticides. This aspect of the turfgrass plant is expected to compete with the transfer of pesticide residues to exposed hands, legs, etc. and reduce dermal exposure levels. The next most significant exposure route involves inhal­ ation of volatile pesticide residues or residues associated with particulate, such as aerosols and dust particles, by the lung during breathing. Although usually not considered as significant as dermal exposures on a quantitative basis, the respiratory route of exposure can be extremely toxicol­ ogically relevant due to the high absorptive rate in the lung and its direct uptake into systemic circulation with little or no involvement with xenobiotic metabolism. The oral route of exposure via the gastrointestinal tract is considered the least extensive and occurs primarily due to hand-to-mouth contact. This exposure route may have more relevancy in children than in adults, although its significance even in children is not well quantified. (a)  Exposure and Hazard from Volatile and Dislodgeable Foliar Residues The hazard associated with the inhalation of volatile pesti­ cide residues and pesticides associated with aerosols and dusts has been indirectly determined by various methods

1063

following application to turf. Most common means involve the use of high-volume air samplers, breathing zone estimates using personal low-volume air samplers, and a variety of small-plot techniques. Dislodgeable foliar residues, likewise, have been determined by various means – including organic solvent- or surfactant-based extractions, surface wipes using dampened gauze or cheesecloth, and clothcovered weighted sleds, rollers, and shoes – and used to estimate dermal exposures and associated hazard. Volatilization may be defined as the loss of chemicals from surfaces in the vapor phase with the subsequent movement into the atmosphere (Spencer and Cliath, 1990). Postapplication vaporization of pesticide residues was reported as early as 1946 (Staten, 1946), and numerous studies since then have established volatilization as a major avenue of pesticide loss following application (Taylor, 1978). Although pesticide volatilization following soil application has been fairly well documented, volatilization from plant canopies has been studied far less (Cooper, 1993). A dense perennial groundcover, such as turfgrass, is quite different from a plowed field or field crop planting and is likely to provide a unique environment for volatilization. The high surface area associated with the dense turfgrass blades and the thatch layer, which retards the downward leaching of pesticides, is likely to result in a significant increase in volatile and dislodgeable pesticide residues (Murphy et al., 1996a,b). Understanding the nature and magnitude of volatilization and dislodgeable foliar residues is important not only because of its impact on pesti­ cide dissipation but also because of concerns regarding pesticide efficacy and human exposure via inhalation and dermal penetration, respectively. Unfortunately, research evaluating pesticide volatilization and dislodgeable foliar residues from turfgrass has been quite limited. Volatilization of pesticide residues from plants under field conditions often exhibits marked diurnal fluctuations with maximum loss occurring at approximately solar noon (Cooper et al., 1990; Taylor et al., 1976, 1977). This diurnal variation is driven primarily by solar heating. During mid-afternoon, solar heating is at a maximum, resulting in elevated surface temperatures and increased atmospheric turbulence. Prior to sunrise and after sunset, however, there is little insolation to elevate surface temperatures, thus resulting in minimal volatility during these periods (Murphy et al., 1996a,b). Loss of pesticides by volatilization from foliage typic­ ally follows a diphasic decline with an initial rapid loss for approximately 1 week, followed by a period of much slower volatile loss (Spencer et al., 1973; Taylor et al., 1977). The slower rate of volatile loss typically observed after the first week may be explained by two hypotheses (Taylor, 1978). The first suggests that the remaining residues are less available because they lie deeper within the canopy and are trapped in irregular areas of leaves, stems, and leaf–stem junctions. The second suggests that the latter residues are lost at a reduced rate because they are more strongly

Hayes’ Handbook of Pesticide Toxicology

1064

adsorbed or have penetrated the leaf surface. Both mech­ anisms contribute to volatile loss and the availability of dislodgeable foliar residues, but the relative importance of each factor has not been determined (Murphy et al., 1996a,b). The initial experiments that first examined the volatile and foliar loss of pesticides following application to turfgrass were described in a series of papers (Jenkins et al., 1990, 1991, 1993). Airborne and dislodgeable residues were determined following application of the dinitroaniline herbicide pendimethalin to Kentucky bluegrass. In the first technique, airborne residues were measured using small field chambers consisting of 19-liter Pyrex bottles without bottoms and fitted with Teflon cartridges containing XAD-4 polymeric resin. Dislodgeable foliar residues were determined from turfgrass plugs in which the grass blades were separated from soil and thatch and then extracted with methanol. The focus of this initial research was to determine what relationship, if any, exists between dislodgeable foliar residues and pesticide volatility. If such a relationship exists, a mathematical model could be developed to estimate pesticide flux using only measurements of dislodgeable foliar residues, thereby eliminating the logistical and conceptual problems associated with estimating volatile loss using traditional micrometeorological or laboratory chamber methods. Also, the relationships between volatile loss and the physical parameters of temperature/solar radiation and wind speed could be investigated in a controlled and replicated manner. These studies found that it was possible to estimate total daily flux based on the diurnal pattern of volatility by making single flux measurements during the period of peak volatility (1300–1500 h). These authors concluded that there was a relationship between volatile loss and dislodgeable foliar residues because both flux and dislodgeable foliar residues exhibited similar biphasic patterns of decline over the course of the study. However, the measured values did not correlate well, suggesting that the relationship was more complicated than initially thought. Although the field chambers allowed the examination of different variables in the attenuation of volatile and dislodgeable residues, the modification of the chamber environment with regard to temperature and wind speed made it unclear whether the measured values of “chamber flux” actually reflected what was happening outside the chamber. It was concluded that the effects of increased temperature and decreased wind speed in the chambers were probably offsetting each other in this case, but that the chamber technique could not reasonably be expected to reflect environmental volatility in all cases. As such, the field chamber technique is not appropriate for studies that require an accurate determination of ambient pesticide air concentration (e.g., exposure studies) as opposed to an estimate of pesticide flux. A second group of experiments was conducted to compare flux rates from the chambers with those obtained from an ambient method, the theoretical profile shape (TPS)

method, which is based on the trajectory-simulation (TS) model of Wilson and co-workers (Majewski et al., 1989, 1990, 1991; Wilson et al., 1982, 1983). Briefly, the TPS method employs the two-dimensional TS dispersion model to estimate source strength [Fz(0)] from a single measurement within the vertical profile of the horizontal flux at the center of a circular plot. The model can be written as

Fz (0)  (uc)measured / 



where Fz(0) is the source strength determined as the actual vertical flux rate (mg/m2/h), (uc)measured is the product of the measured wind speed and air concentration, and  is the normalized horizontal flux predicted by the TS model. The measurement height (ZINST) is chosen based on the plot radius, roughness length (z0), and the Monin– Obukhov atmospheric stability length (L). In these experiments, the airborne residue and wind speed were measured at a height ZINST of 73 cm chosen according to a plot radius of 20 m and a surface roughness length of 0.2 cm. Pendimethalin airborne residues were collected with a Staplex TF1A high-volume air sampler onto XAD-4 polymeric resin. Dislodgeable foliar residues were determined as described for the field chamber experiments. Previous work by Majewski and co-workers (1989, 1990, 1991) determined the TPS method to be comparable to other micrometeorological techniques for estimating evaporative flux of a number of pesticides from soil. In the case of pesticide application to turfgrass, the two smallplot techniques gave comparable results for pendimethalin flux. The important difference between the two methods is that although the TPS method has inherent error associated with the mathematical TS model used to calculate flux, its results are based on actual air concentrations, whereas the levels measured in the field chambers reflect an artificial environment and ultimately require verification with an ambient technique. In addition, the TPS method provides an accurate measure of pesticide air concentration, which can be used for purposes other than estimating flux, such as exposure estimates and hazard evaluations. The TPS method, likewise, uses plots small enough to allow for replication, which is a major limitation in the evaluation of information gathered using aerodynamic-based methods. Using a small circular plot design in conjunction with the TPS method, airborne pesticide concentrations were determined and used to estimate the inhalation and dermal exposure situations for golfers using the U.S. EPA hazard quotient determination (Clark et al., 2000; Murphy et al., 1996a, 1996b). An average daily inhaled dose of pesticide for a 70-kg adult playing a 4-h round of golf was estimated as follows

C  R  4 h / 70 kg  Di

(1)

where C is the measured air concentration of pesticide determined by high-volume air sampling (g m3), R is

Chapter | 47  Lawn and Turf: Management and Environmental Issues of Turfgrass Pesticides

the adult breathing rate during moderated activity (2.5 m3 h1; U.S. EPA, 1989), and Di is the daily inhaled dose of pesticide (g kg1). The estimated inhaled dose (Di) is divided by a chronic reference dose (RFD, g kg1 day1; U.S. EPA, 1993), resulting in the inhalation hazard quotient (Di/RFD  IHQ). Similarly, an average 4-h dermal dose was calculated using

S  P / 70 kg  1000 mg mg1  Dd

(2)

where S is the calculated dermal exposure (mg), P is the average estimated dermal permeability (0.1), and Dd is the dermal dose of pesticide (g kg1). S is determined by multiplying the maximum dislodgeable foliar residues determined from a cheesecloth wipe obtained during the first day of sampling (Thompson et al., 1984) by a dermal transfer coefficient of 5  103 cm2 h1 (Zweig et al., 1985) times 0.25 to scale down the interaction with the medium from that which would likely occur with a golfer compared to a strawberry harvester times 4 h. The estimated dermal dose (Dd) is divided by the chronic RFD, resulting in the dermal hazard quotient (Dd/RFD  DHQ). Chronic RFDs were determined from daily doses shown to cause no observable effects on laboratory animals over their lifetime and then divided by safety factors of 10 to 10,000, depending on the completeness of the toxicological data set (U.S. EPA, 1993). Thus, RFDs are deemed to be safe doses that can be received over a lifetime without causing adverse effect. From this, HQs less than or equal to 1.0 indicate that the residues present are at concentrations below those that could cause adverse effects in humans. An HQ value greater than 1.0 does not necessarily infer the residue levels will cause adverse effects but, rather, that the absence of adverse effects is less certain. Using the preceding format, four pesticides were initially examined: two insecticides, isazofos and trichlorfon; a herbicide, MCPP; and a fungicide, triadimefon. The hydrolytic degradative product of trichlorfon, DDVP, was also determined. Only the application of isazofos resulted in both IHQs and DHQs greater than 1.0 over the course of this study. In addition, the hydrolytic degradative product of trichlorfon, DDVP, also resulted in a DHQ value greater than 1.0, but this only occurred in the combination with 1.3 cm of postapplication irrigation. Thus, it is apparent that exposure situations exist following application of pesticides to turfgrass that cannot be deemed completely safe using the U.S. EPA HQ criteria. Furthermore, the pesticides of most concern appear to be insecticides that have inherently high vapor pressure (high volatility) and relatively high inherent toxicity (low RFDs). To validate these findings, a larger study was conducted that included 13 pesticides (plus DDVP) used extensively on turfgrass, which varied widely in terms of their vapor pressure, water solubility, and inherent toxicity (Clark et al.,

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2000). Of the 14 pesticides examined, only three (ethoprop, diazinon, and isazofos) resulted in volatile residues at sufficiently high concentrations to produce IHQs greater than 1.0 over the entire time course of this study (Table 47.6). As in the initial study, all three pesticides were organophosphorous insecticides that have high vapor pressures (high volatility) and low RFDs (high inherent toxicity). As indicated in Table 47.7, seven pesticides resulted in dislodgeable foliar residues at sufficiently high concentrations to produce DHQs greater than 1.0. Except for bendiocarb (a carbamate), all the rest were organophosphorous insecticides. Only ethoprop, isazofos, diazinon, and isofenphos produced DHQs that exceeded 1.0 in the intervals beyond the first 24-h period following application. Trichlorfon, chlorpyrifos, and bendiocarb had DHQs only slightly above 1.0, and these fell below 1.0 after the first day following application. These results are consistent with the original findings and substantiate that there are exposure situations involving volatile and dislodgeable foliar residues following the application of selective pesticides (organophosphorous insecticides) to turfgrass that cannot be deemed completely safe using the U.S. EPA HQ criteria. In addition, increased hazard appears to be well correlated with pesticides that have high vapor pressures and low RFD values, and these characteristics may be useful in predicting hazard associated with other related pesticides not included in the present study. With this approach, selected pesticides, which possess high volatility and toxicity, may result in exposure situations that cannot be deemed completely safe as judged by the U.S. EPA HQ criteria. This assessment, however, must be viewed in terms of the assumptions that were used in making these estimations. In all instances, maximum pesticide concentrations were used for the entire 4-h exposure period, maximum rates for pesticide applications were used, and dermal transfer coefficients and dermal penetration factors were taken from non-turfgrass situations that are likely to exceed those that would take place on a golf course. Thus, such estimates are worst case scenarios. To more accurately predict the health implications of pesticide exposure on golfers and during other turfgrass activities, a relevant dosimetry and biomonitoring evaluation needs to be carried out. With more direct exposure estimates, the previously reported exposure levels are likely to be in excess of the true exposure to pesticides on a golf course and on managed turfgrass. Initial work with this more direct approach was conducted by Yeary and Leonard (1993), who measured pesti­ cides in air during and after application to lawns, trees, and shrubs in urban environments. Air sampling was conducted according to procedures given in the Occupational Safety and Health Administration’s Industrial Hygiene Technical Manual using low-volume personal air samplers and trapping pesticides onto charcoal resin. The pesticides measured were acephate, atrazine, carbaryl, chlorpyrifos, 2,4-D, diazinon, dicofol, PCPA, and pendimethalin.

Table 47.6  Inhalation Hazard Quotientsa for Turfgrass Pesticides in the High (Vapor Pressures 1.0    105 mm Hg), Intermediate (Vapor Pressures between 1.0    105 and 1.0    107 mm Hg), and Low (Vapor Pressures 1.0    107 mm Hg) Vapor Groups Pesticideb

Vapor pressure (mm Hg)

OPP RFD (mg/kg/day)

IHQs Day 1

Day 2

Day 3

0.02 1.2 1.2 3.4 0.04

High vapor pressure          

DDVP Ethoprop Diazinon Isazofos Chlorpyrifos

1.6  102 3.5  104 9.0  105 5.6  105 2.0  105

5.0  104 1.5  106 9.0  105 2.0  105 3.0  103

0.06 50 3.3 8.6 0.09

0.04 26 2.4 6.7 0.1

3.8  106 3.4  106 3.3  106 5.7  107 4.2  107 3.1  106

2.0  103 5.0  103 5.0  104 1.5  103 1.25  102 1.4  102

0.02 0.02 n/d 0.001 n/d 0.0005

0.004 0.002 0.02 0.001 n/d 0.0001

0.004 0.002 n/d 0.0003 n/d 0.00004

7.1  108 3.8  109 2.0  109

8.0  102 6.1  102 2.5  102

n/d n/d n/d

n/d n/d n/d

n/d n/d n/d

Intermediate vapor pressure            

Trichlorofon Bendiocarb Isofenphos Chlorothalonil Propiconizole Carbaryl

Low vapor pressure   Thiophanate-methyl   Iprodione   Cyfluthrin a

 The IHQs reported are the maximum daily IHQs measured, all of which occurred during the 11:00 am to 3:00 pm sampling period.  All pesticides were watered in following application using 0.63-cm postapplication irrigation. IHQ, inhalation hazard quotient; n/d, not detected; OPP RFD, Office of Pesticides Program (U.S. EPA) reference dose.Adapted from Clark, et al. (2000), Table 5.6, p. 305. b

Table 47.7  Dermal Hazard Quotients for Turfgrass Pesticides Listed with Increasing RFDs from Top to Bottom through Day 3 Postapplication Pesticidea

RFD (mg/kg/day)

Day 1 (DHQs)

Day 2 (DHQs)

Day 3 (DHQs)

5 Hours

8 Hours

12:00 pm

12:00 pm

Ethoprop

0.000015

190

156

26

39

Isazofos

0.00002

135

112

18

24

Diazinon

0.00009

32

25

Isofenphos

0.0005

5.7

5.7

DDVP

0.0005

0.3

0.3

Trichlorofon

0.002

0.8

1.1

0.9

0.5

Chlorpyrifos

0.003

2.3

1.8

0.3

0.4

Bendiocarb

0.005

0.7

1.1

0.7

0.09

Propiconizole

0.0125

0.3

0.02

0.05

0.02

Carbaryl

0.014

0.009

0.01

0.007

0.0002

Cyfluthrin

0.025

Ipridione

0.061

Thiophanate-methyl

0.08

a 

n/a 0.03 n/a

n/a 0.03 n/a

4.6

5.7

1.1

1.1

n/d

n/a 0.04 n/a

n/d

n/a 0.03 n/a

All pesticides were watered in immediately following application with 0.63 cm of postapplication irrigation. DHQ, dermal hazard quotient; n/a, not available; n/d, not detected; RFD, reference dose.Adapted from Clark et al. (2000), Table 5.7, p. 306.

Chapter | 47  Lawn and Turf: Management and Environmental Issues of Turfgrass Pesticides

Air samples were taken for breathing zone air of pesticide applicators, indoor air of residential properties, and ambient air of residential properties. Approximately 500 samples were taken in 14 cites in the United States and Canada. Monitoring results indicated that 80% of the samples were below the detection level of 0.001 mg/m3. When analytes were detected, the time-weighted average concentration was less than 10% of any established standard for occupational exposure. Dislodgeable foliar residues following the application of pesticides to turfgrass have also been evaluated and vary depending on the collection method used and the type of formulation applied. Thompson et al. (1984) first reported on the persistence and dislodgeable residues of 2,4-D following application to Kentucky bluegrass. Sprayable liquid formulations had higher dislodgeable foliar residues compared with a granular product immediately after application. Residues rapidly dissipated after application, with less than 1% of the total applied 2,4-D being available as dislodgeable residues as determined by dampened cheesecloth wipes. Irrigation or rainfall after application and mowing significantly reduced dislodgeable residues. Bowhey et al. (1987) found that dislodgeable residues of 2,4-D increased proportionally as the rate of application was increased but decreased rapidly with less than 1% of the total applied remaining 7 days following application. Dislodgeable residues immediately following application were greater for liquid formulations than for granular products but returned to equality by 24 h following application. Goh et al. (1986) studied dislodgeable foliar residues of chlorpyrifos following application to Kentucky bluegrass using a surfactant-stripping procedure. Using this method, dislodgeable residues of chlorpyrifos were determined to be below the estimated safe level of 0.005 mg/cm2. Irrigation immediately following application significantly increased the dissipation rate of chlorpyrifos residues from foliage. In a laboratory study, Sears et al. (1987) found that dislodgeable foliar residues of diazinon removed immediately following application by mechanical wiping with dampened cheesecloth constituted approximately 10% of the total amount applied but declined to less than 0.3% within 1 day. Cowell et al. (1993) compared several techniques for determining dislodgeable foliar residues of turf pesticides and found that wiping with polyurethane foam pads resulted in good estimation of reentry exposure to herbicides. Polyurethane foam has a high affinity for lipophilic organic chemicals and residues remain stabile for storage. In a practical reentry situation, only a small portion of the treated foliage is expected to be contacted, and skin and clothing materials are expected to have less affinity for turf pesticides than the polyurethane foam pads. Thus, this technique is expected to yield worst case estimates of the potential for dermal exposure from turf-applied pesticides.

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Black and Fenske (1996) compared water-­dampened gauze pad wipes and water/surfactant foliar washes to estimate the dislodgeability of chlorpyrifos. The wipe technique removed an average of 0.50 g/cm2 of chlorpyrifos, or 1.5% of the initial deposited residue. The foliar wash technique recovered significantly more chlorpyrifos (1.3 g/cm2, or 4.0% of total applied). Results with the wipe technique also indicated that residues obtained from irrigated plots were significantly lower than from nonirrigated plots during the first 6.5 h following application. Foliar washes resulted in dislodgeable residue levels that were significantly lower from irrigated compare with nonirrigated plots after the first 3.5 h following application. These results indicate that foliar washes remove larger amounts of chlorpyrifos from turf than the wipe technique and would therefore yield higher estimates of potential exposure if used as a turf contact-transfer factor. (b)  Exposure Determinations Using Dosimetry and Biomonitoring for Applicators Initial efforts to determine direct exposure following application of turf pesticides were carried out by Weisskopf et al. (1988), who studied the personnel exposure to diazinon in a supervised pest eradication program for larvae and adult Japanese beetles. Dermal and respiratory exposure was assessed for 15 workers (supervisors, crew chiefs, and applicators) applying a granular diazinon formulation to lawn turf environments by standard industrial procedures. Whole-body dosimetry was determined using batterypowered personal air pumps, ethyl alcohol hand washes, and gauze sponge patches. Biomonitoring was determined by the analysis of urine for the diazinon metabolite DETP (O,O-diethyl ester of phosphorothioic acid). Diazinon exposures ranged from 0.1 to 11 mg/day for applicators and from 0.03 to 0.3 mg/day for supervisory personnel over a 3.3- to 7-h sampling period. Exposure for workers correlated with job classification and practices, with the magnitude of exposure in the following descending order; belly grinder users, regular crew members, crew chiefs, and supervisors. Lawn care professionals were monitored for exposure to dithiopyr, a turf herbicide, by both passive dosimetry and biomonitoring techniques during work (Cowell et al., 1991). Passive dosimetry was performed using cotton gauze patches, silica gel air sampling, and hand washes. Biomonitoring was determined by the analysis of 72-h urine samples for the dithiopyr metabolite, dicarbothioic acid. The mean body dose estimate from biomonitoring of urine samples was 4.6  105 mg/kg/lb applied. The passive dosimetry body dose estimates were corrected for skin penetration and calculated for two clothing scenarios. The mean body dose estimate for a fully clothed professional wearing a long-sleeved shirt was 9.1 mg  105 mg/kg/lb compared to an individual wearing a short-sleeved shirt who was exposed to 3.6  104 mg/kg/lb. The lower leg

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regions had the highest exposure as determined by passive dosimetry.

47.3.2.2  Wildlife and Other Nontargets The potential for adverse impacts on wildlife and other nontarget organisms, including birds, wild mammals, amphibians, soil invertebrates, and microorganisms, to turf pesticides is high and merits consideration. Urban turf environments and golf courses represent significant wildlife habitats, particularly for birds (Balogh, 1992; Brewer et al., 1993; Hooper et al., 1990). The relative high frequency and abundance of birds in the urban lawn environment are indicative of the potential for exposure to turf pesticides, but surveys of activity and habitat use provide more relevant information on the relative potential of exposure among various species (Brewer et al., 1993). Thus, turf-dwelling birds such as the American robin, common grackle, and the European starling, which probe thatch and upper soil layers for food, are at greater risk of exposure to turf pesticides from both direct contact and ingestion of contaminated food. It is also noteworthy that wildlife habitats on and surrounding golf courses usually are not routinely treated with pesticides, and the most heavily treated aspects (tees and greens) tend to be of limited value for wildlife. Detailed information on bird intoxication resulting from organophosphorous and carbamate insecticides used on turf is not abundant in the literature. Grue et al. (1983) authored one of the first accounts of the hazards associated with organophosphorous insecticides on wildlife, and Stone (1979) and Stone and Gradoni (1985) reported numerous wildlife kills resulting from the use of organophosphorous insecticides on turf. Zinkl et al. (1978) reported a die-off of Canada geese following an application of diazinon to turf, and Kendall et al. (1992) documented diazinon-related mortality of American wigeons on turf. Dietary exposure studies of European starlings on urban lawns indicated a clear preference for insect larvae from the family Noctuidae (Brewer et al., 1993). Monitoring of insecticide residues in larvae established that these insects do accumulate organophosphorous insecticides and therefore constitute an exposure route to birds that eat them for food. These authors additionally established a food chain transport of the residues by collecting fecal–urine samples from starling nestlings that were contaminated by these insecticides. However, it is important to note that the establishment of exposure does not necessarily indicate that intoxication will occur. Brewer et al. (1987) and Kendall et al. (1993) found that low-level diazinon exposure in passerine species and in Canada geese resulted in no mortality or other observable adverse effects. In an extensive study of avian exposure to organophosphorous and carbamate pesticides on a coastal South Carolina golf course, Rainwater et al. (1995) reported a clear potential for exposure to the turf pesticides, but few actual exposures and acute toxic effects were observed.

Hayes’ Handbook of Pesticide Toxicology

Research on the effect of pesticides on soil invertebrates and earthworms in turf environments indicates that their potential for causing adverse effects is very chemical-class specific (Potter et al., 1993, 1998). Specifically, the use of most herbicides and some insecticides (2,4-D, dicamba, and isofenphos) resulted in little, if any, adverse impact on earthworm populations, whereas some insecticides (ethoprop, carbaryl, and bendiocarb) and fungicides (benomyl) caused severe toxicity. The abundance of Cryptosigmata, Collembola, and ants has also been determined to be reduced drastically by some insecticides. These pesticides additionally have been shown to reduce the rate at which earthworms incorporate mineral soil into submerged thatch. Similarly, certain fungicides may reduce soil pH, which decreases the ability of microorganisms to decompose thatch (Smiley and Fowler, 1986). In addition, some fungicides have increased the rates of root and rhizome production and can contribute to enhanced thatch accumulation (Smiley et al., 1985). Although there is substantial research that establishes that the use of certain turf pesticides can adversely affect soil invertebrates and thatch degradation, there is information that suggests that the impact of a well-managed, high-maintenance lawn care program is less severe than indicated previously (Arnold and Potter, 1987). In these experiments, replicated plots of Kentucky bluegrass were maintained for 4 years on a maintenance schedule typical of that used by many professional lawn care companies. At the end of the 4-year experiment, earthworm populations were not significantly reduced, the number of oribatid mites increased, and soil pH only decreased slightly. Thatch accumulation was significantly greater but not excessive. Predator populations (Araneae, Staphylinidae, and Carabidae) were significantly decreased but rebounded the next year after the maintenance program had ended. As with many agricultural situations, the indiscriminant, excessive, and repeated application of pesticides can lead to acquired resistance in pest populations and to increased microbial degradation of turf pesticides. Resistance to organophosphorous and carbamate insecticides is well documented for chinch bugs, greenbugs (Reinert, 1982), and white grubs (Ahmad and Ng, 1981). With the rapid development of resistance in virtually every highly managed plant system, the general adoption of resistance management protocols for turf is wise if not necessary. Resistance management strategies include: reducing the frequency and extent of pesticide application; limiting the use of pesticides with extensive environmental persistence and those in slow-release formulations; reducing the use of residual treatments; avoidance of treatments that select against both immature and adult stages; and increased use of cultural, biological, and other nonchemical methods used in IPM approaches (Metcalf, 1989). The enhancement of microbial degradation of pesticides following repeated applications of the same or similar pesticides in soil environments is well documented in

Chapter | 47  Lawn and Turf: Management and Environmental Issues of Turfgrass Pesticides

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a variety of agricultural settings and has been reported in turf for isofenphos, diazinon, ethoprop, etc. (Felsot, 1989; Niemczyk and Chapman, 1987). It was also shown that carbaryl and diazinon degraded more rapidly in thatch taken from plots previously treated with isofenphos. Similarly, thatch from turf previously treated with diazinon, chlopyrifos, carbaryl, or isazophos showed enhanced degradation rates for these pesticides.

process also helps to evaluate which management practices work effectively. 6. Education to develop ongoing employee training focusing on pest recognition and agronomic factors that help them to make sound management decisions. Client education on management action to reduce pest problems (soil tests, aerification, dethatching, fertilization, and tree care) while providing a safe, environmentally responsible turf facility.

47.3.3  Best Management Practices

Through research sponsored by the USGA, a number of projects produced results that will help with one or more of the steps outlined previously in expanding the utility of turf IPM practices. For example, in North Carolina, with cooperation of Cornell University, extensive research was conducted on mole cricket behavior (Brandenburg and Villani, 1998). The results indicate that mole crickets are capable of avoiding pesticides applied to control them by burrowing deep into the soil until the active ingredient has degraded. A better understanding of mole cricket response to pesticide applications will help determine proper placement and timing of products that enhance contact with the insect. Research conducted on white grubs (scarab beetle larvae) in Kentucky examined the role of several cultural practices on insect outbreaks (Potter, 1998). Field studies showed that withholding irrigation during peak flights of beetles, raising cutting height, and a light application of aluminum sulfate in the spring helped to reduce the severity of subsequent infestations of Japanese beetle and masked chafer grubs. Grub densities were not affected by spring applications of lime or urea, but use of organic fertilizers (composted cow manure or activated sewage sludge) increased problems with green June beetle grubs. Use of a heavy roller was not effective for curative grub control. Soil moisture was the overriding factor determining distributions of root-feeding grubs in turf.

A significant conglomeration of research findings related to the potential impacts and behavior of turf pesticides needs to be amassed to ensure that the most effective and safe use of such chemicals will occur. This body of information includes development and registration information necessary to introduce a new product commercially, which is usually supplied by the registrant; product stewardshipbased research to uphold the highest standards of ethics and environmental responsibility by addressing issues on product quality, performance, and efficacy; and independent research that complements the more proprietary industrial approaches, which is usually completed at universities, federal or state laboratories, or private research organizations (Racke, 2000). This information is used by regulators, manufacturers, researchers, consultants, and lawn care professionals in making informed pest management and risk management decisions. Following are ex­amples of research-generated approaches that form some of the best strategies for turf management.

47.3.3.1  Application of Integrated Pest Management This approach relies on a combination of preventative and corrective measures to keep pest densities below levels that would cause unacceptable turf damage. Its goal is to manage pests effectively, economically, and with minimal risks to people and the environment. The process of IPM involves the following steps and has resulted in a number of useful research findings, as summarized by Kenna and Snow (2000): 1. Sampling and monitoring turf areas to detect pests and evaluate how well control tactics have worked. 2. Pest identification to adequately understand the habits and life cycle of the organism and how it can be managed. 3. Decision making guided by action thresholds based on pest damage that justifies treatment or intervention. 4. Appropriate intervention that determines why a pest outbreak occurred and if cultural practice adjustments can reduce damage or future outbreaks. 5. Follow-up that includes recordkeeping to help predict future pest problems and to plan accordingly. This

47.3.3.2  Modeling and Risk Assessment The perceived risks and the actual risks to humans, other nontarget organisms, and the environment that are attributed to pesticide use on turf are in most cases not similar and in many cases err in an overestimation of the risks involved (Durborow, 2000). The environmental and exposure issues of turf management summarized previously in this chapter are limited but encouraging. A means of extrapolating such findings from one turf environment to another is computer simulation modeling. The use of computer simulation models to predict chemical impact at the field and watershed scale is well documented for many agricultural settings and is currently being established for turf environments (Cohen et al., 1993).The most rigorous means to conduct quantitative risk assessments for pesticide leaching and runoff potential is by the use of computer simulated modeling. Nevertheless, existing models produce

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results that are inexact. The following sections summarize the current status of modeling pesticide movement in the turf environment. (a)  Leaching and Runoff One of the first modeling exercises for leaching via turf environments was initiated by Smith et al. (1991), who used the pesticide root zone model (PRZM) and the groundwater loading effects from agricultural management systems (GLEAMS) model to simulate the leaching and movement of atrazine and alachlor herbicides in Bahia grass turf plots in Georgia. Both models performed well in the prediction of time-to-peak concentrations of these herbicides at 0.61 (2 ft) and 1.22 m (4 ft) depths and resulted in fair predictions of peak concentration. Several computer models have been used to predict surface runoff losses of water, sediment, and chemicals including pesticides; these include the GLEAMS and the simulator for water resources in rural basins, water quality (SWRRBWQ) models. The SWRRBWQ model was used by Cohen et al. (1993) to estimate the extent of runoff and leaching for 21 pesticides from a proposed 300to 400-acre golf course in a forested area in Hawaii. The SWRRBWQ model predicted the runoff of the herbicide 2,4-D to be 0.03% of the total applied and 1.83% for the fungicide mancozeb. Interestingly, Squillace and Thurman (1992) determined from a review of the literature on the runoff of herbicides from agricultural environments and watersheds that this value typically ranges from 0.3 to 5% of the total amount applied. Thus, these models appear able to make close approximations of runoff concentration of pesticides and apparently are consistent with the fact that runoff is lower in turf environments than in agricultural cropping situations, such as corn and soybeans. From this information, alternative pesticides with minimal potential for aquatic impacts were recommended. A computer simulation using the SWPPBWQ model of a Florida golf course in conjunction with a monitoring study was performed by Cohen et al. (1993). Runoff water was sampled during selected storm events and analyzed for basic water quality parameters, including pesticides. It was found that the model failed to predict certain water and sediment runoff events, most likely due to the low amounts of water and sediment erosion that occurred on the relatively flat and sandy turf environment that existed on the golf course. Using the PRZM model coupled with the Vadose zone flow and transport (VADOFT) model, Cohen et al. (1993) predicted the leaching potential for 2,4-D and mancozeb between two Hawaiian golf course sites that differed significantly in the amount of rainfall that each received. Soil water concentrations of both pesticides were at least one order of magnitude greater for the wet site compared with the dry site. The estimated concentrations at either site, however, were considerable less than their respective HALs.

Hayes’ Handbook of Pesticide Toxicology

Smith and Tillotson (1993) also employed the GLEAMS model to predict the potential for herbicides to leach through simulated Bermuda grass and bentgrass greens and to compare the predictions with actual lysimeter-based experimental data. Although only minute quantities of herbicides (5 ppb) were detected in the actual lysate, the GLEAMS model significantly overpredicted the herbicide concentrations in all cases. One possibility for the differences between the measured and predicted concentrations of leached herbicides in this study was identified as an incomplete understanding of the quantitative fate of the herbicides on the vegetative surface of the turfgrass and the role that the thatch layer plays in herbicide transport. The probable reason for no significant influence by the thatch layer is the thin thatch layer estimate (1.4 mm) used in the model and the limited increase in organic matter used by the model compared to that actually measured in the rooting mixture (5.8 vs. 2.26% w/w). The transport of the postemergence herbicide dicamba in turfgrass thatch and foliage has been studied and modeled (Carroll et al., 2000). The retention of dicamba by turfgrass foliage and thatch and the transport of dicamba in undisturbed columns of bare soil and turfed soil were evaluated using the linear equilibrium (LEM) and the twosite nonequilibrium (2SNE) models. The 2SNE model gave reasonable estimates of dicamba transport when the model retardation factor (R) was calculated using laboratory-derived adsorption coefficients. The LEM model satisfactorily predicted dicamba transport only when R was curve fitted. The use of soil column retardation factors that were based on laboratory-derived adsorption coefficients resulted in poor estimates of dicamba by the LEM model. A novel approach in predicting pesticide and nitrogen leaching in turf environments has been developed using artificial neural network (ANN) modeling techniques (Starrett and Starrett, 2000). The major difference between traditional approaches and the ANN technique is that in the traditional method, the modeler creates mathematical relationships between inputs and outputs, whereas the ANN program learns the relationships based on the training data set. Traditional models, such as GLEAMS, PRZM, and SWRRBWQ, are created by starting with assumed mathematical relationships between the variables under consideration. The success of the models depends on how well the theoretical equations match the real behavior of the system being evaluated. The ANN technique directly utilize data measured in the real system to actually create the model. Thus, the ANN technique forms a functional relationship between the given inputs and the desired outputs without the need of a formal theory or mathematical model. Using this approach, these researches have developed two models: the pesticide ANN model, which requires only the pesticide solubility, Koc, time after application, and the irrigation practice; and the nitrogen ANN model, which requires the nitrogen form, percentage of the soil that is

Chapter | 47  Lawn and Turf: Management and Environmental Issues of Turfgrass Pesticides

sand, time after application, and the irrigation practice, to “train” the model. Most of the training and testing data were derived from leaching studies that used 50-cm turfed undisturbed soil columns. Overall, the predictive ability of these models was within approximately 4% of the test case values, indicating the feasibility of this approach. (b)  Volatilization Modeling approaches have also been attempted for the assessment of potential inhalation exposures from volatile pesticides applied to turf (Haith et al., 2000). Vapor losses of pesticides from turfgrass are influenced by chemical properties, environmental and site conditions, and management practices. Estimates of the potential concentration inhaled by humans involved in turf activities are required for the assessment of possible health risks. Mathematical models can potentially be used to estimate volatilization of pesticides applied to turf, but their complexity and limited testing have restricted their general applicability to the realities of field conditions. The simpler empirical approach of volatilization indices may be a more feasible option. Measured pesticide concentration can be directly related to basic chemical properties and environmental variables, and these same relationships can be used in an evaluation to identify hazardous situations. Volatilization indices of eight pesticides used on turf (bendiocarb, carbaryl, chlorpyrifos, diazinon, ethoprop, isazofos, isofenphos, and trichlorfon), based on their application rates, Henry constants, vapor pressures, organic carbon adsorption coefficients, and wind speed, show considerable promise as practical tools for exposure assessment. The regressions of three of these indices (vapor pressure, volatilized mass divided by air velocity, and volatilized mass calculated with vapor pressure replacing the Henry constant divided by air velocity) explained 70–90% of the variation in measured air concentrations of the eight test pesticides applied to turf plots. The classification of these pesticides using the U.S. EPA IHQ criteria were identical whether the classification was made using actual measured air concentrations or with concentrations estimated by the regression of any of these three indices.

47.3.3.3  Operational and Cultural Practices (a)  Postapplication Irrigation The watering-in of pesticides immediately following their application (postapplication irrigation) to turfgrass has long been a suggested and sometimes required practice to ensure efficacious pest control and to minimize dermal and inhalation exposures upon reentry. A quantitative assessment of the effects of postapplication irrigation has been made and indicates the need for more carefully managed irrigation practices for turf (Clark et al., 2000; Murphy et al., 1996a,b). Using a 1.3-cm level of postapplication irrigation, two unwanted processes occurred concerning the fate and

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availability of selected turf pesticides. First, trichlorfon was determined to be converted into its more toxic, volatile, and water-soluble metabolite, DDVP, at a higher level than with a 0.63-cm level of irrigation. Second, the higher level of postapplication irrigation appeared to delay the appearance of the maximal level of volatile and dislodgeable foliar pesticide residues from the day of application (at 0.63 cm) to days 2 and 3 following application (at 1.3 cm). This delay is of particular concern because it places an increasing number of individuals involved in turf activities on the turf at a time when exposure is most likely. To demonstrate this aspect independently of the conflicting factor of different application regimes, isazofos was applied in an identical manner and the two different levels of postapplication irrigation were directly compared (Clark et al., 2000). To illustrate the dramatic effect that postapplication irrigation has on the initial levels of dislodgeable foliar residues, samples were collected 15 min following application but before irrigation. Both levels of irrigation greatly reduced the initial levels of dislodgeable foliar residues, but the 1.3-cm level was more effective than the 0.63-cm level. After irrigation, however, the maximal hazard associated with isazofos occurred on day 2 (DHQ  17.1) in the presence of 1.3 cm of irrigation, whereas the maximal hazard in the presence of 0.63 cm of irrigation occurred at 3 h postapplication (DHQ  12). In a related study on USGA-specified greens established with Tifdwarf Bermuda grass, Cisar and Snyder (2000) found that a low irrigation treatment delayed but did not reduce the cumulative leaching of fenamiphos and its more water-soluble metabolites. These results suggest that for fenamiphos and its more mobile metabolites, the use of postapplication irrigation management alone cannot ensure that leaching of these materials will not occur in an envir­ onment such as subtropical south Florida where appreciable rainfall is likely at any time of the year. Nevertheless, irrigation management did reduce the leaching of these pesticides during rainfall restricted periods. These findings indicate that the judicial use of post­ application irrigation in combination with managed spray volume and sprayer configurations may be an effective means to attenuate the hazards associated with exposures to volatile and dislodgeable foliar residues associated with pesticide-treated turfgrass. (b)  Adjuvants, Surfactants, and Absorbents To mitigate the exposure potential of some organophosphorous insecticides that have high vapor pressures and inherent high toxicity, the practical use of spray tank adjuvants and the importance of thatch accumulation on the dissipation of volatile and dislodgeable foliar residues following their application to turf have been assessed (Clark et al., 2000). Two adjuvants, Aqua Gro-L, a nonionic wetting agent/penetrant, and Exhalt 800, an encapsulating spreader/sticker, were examined. Because ethoprop

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consistently resulted in the highest IHQ and DHQ values, it was chosen as a problematic organophosphorous insecticide for study in the combined presence with adjuvants. Isofenphos was applied as a less problematic insecticide. For those applications done using spray tank adjuvants, a 2% v/v concentration of adjuvant was used. Volatile and dislodgeable foliar residues were collected and analyzed, and IHQ and DHQ values were calculated as previously published (Murphy et al., 1996a,b). A paired-plot experimental design was used in these comparisons with one group of plots being prepared on turf established in 1991, which was never aerated or dethatched (mature turf), and the other plots being prepared on turf established in 1995 (new thatch). It was estimated by physical examination that the mature turf plots had a thatch layer that was approximately three times thicker than the more newly established turf plots. To evaluate the effects of thatch accumulation on the dissipation of volatile and dislodgeable foliar residues, applications of ethoprop were made. In the first application, ethoprop was applied simultaneously to a mature turf plot and a new turf plot, both in the absence of any spray tank adjuvant. In the second, ethoprop was applied to a mature turf plot that has been dethatched by vericutting in two directions, and the results were compared with results obtained from this same plot prior to dethatching. In no instance did the addition of adjuvants result in substantial reductions in the amount of volatile and dislodgeable foliar residues of ethoprop following its application as determined by IHQ and DHQ values. In fact, the addition of adjuvants generally resulted in slight increases in the IHQ and DHQ estimations between treatments. Thus, neither of the adjuvants tested appeared to reduce the exposure to volatile and dislodgeable foliar residues following the application of ethoprop or isofenphos as judged by IHQ and DHQ determinations, respectively. Similarly, no substantial or consistent differences were found in levels of volatile or dislodgeable foliar residues following the application of ethoprop or isofenphos to mature versus more recently established turf plots or to thatched versus dethatched turf plots. These preliminary results indicated that thatch management may not be a meaningful approach to mitigating the exposure to volatile and dislodgeable residues following pesticide application to turfgrass. Surfactants are also commonly added to spray tank mixtures to promote water penetration in soils, and they may be particularly useful in moving water through thatch in turf environments. It is well established that thatch is an extremely efficient reservoir for many turf pesticides. In the case of groundwater protection, this aspect of the turf environment is certainly beneficial, but in the case of providing efficacious control for soil-dwelling insect pests or ectotrophic, root-infecting fungi, it may be a serious impediment (Schumann et al., 2000).

Hayes’ Handbook of Pesticide Toxicology

To evaluate the use of a commercially available surfactant “primer” on the efficacy and distribution of a systemic nematicide, fenamiphos was applied to Tifgreen Bermuda grass before, during, and after the surfactant application (Cisar and Snyder, 2000). Fenamiphos itself did not reduce Belonolaimus longicaudtus (sting) nematode populations. Indeed, the use of the surfactant in any combination with fenamiphos failed to result in a significant improvement of fenamiphos control. In addition, there was no difference in the distribution of fenamiphos or its metabolites following any of the treatments with the surfactant. Thus, the commercial surfactant did not improve fenamiphos efficacy or distribution in soil. The effect of a cross-linked phenolic polyether absorbent coated onto silica sand and incorporated into a USGA-type green likewise has been investigated as a means to reduce fenamiphos leaching by Cisar and Snyder (2000). They found a 72% reduction in fenamiphos and a 54% reduction of the more water-soluble metabolite in the leachate collected from the absorbent-sand lysimeter compared with the uncoated-sand lysimeter in the first week of the experiment. The comparative reduction declined with time. At 2 weeks following application, there was no fenamiphos collected from the lysimeter leachate that had the adsorbent sand, and the fenamiphos metabolite was reduced by approximately 90%. By 7 weeks, still no fenamiphos was present in the absorbent-sand lysimeter leachate, and the metabolite was reduced to 24% of the amount found in the unamended lysimeter leachate. Thus, the use of polymer-coated sand incorporated in a USGA green construction did significantly reduce the leaching of fenamiphos and fenamiphos metabolite over a short-term interval. (c)  Use of Barriers for Runoff Containment The effects of increasing vegetation buffer length and mowing height and aerification were evaluated to reduce pesticide and nutrient runoff from turf plots (Baird et al., 2000). All buffer treatments reduced chemical runoff compared with no buffer. There were little differences between buffers of differing lengths. The 3.8-cm buffer mowing height did not significantly reduce runoff losses compared with the 1.3-cm buffer height. However, significant reductions in chemical runoff occurred in the presence of the 7.6-cm buffer height. Aerification of the buffers did not significantly reduce surface runoff losses. From this investigation, the following management practices have been recommended to reduce pesticide and nutrient runoff losses from turf: (1) Establish a vegetation buffer between surface water and treated turf environments; (2) maintain Bermuda grass buffers at a mowing height of at least 7.6 cm; (3) avoid application of turf chemicals when high moisture conditions exist; and (4) develop pest and nutrient management programs that utilize pesticides and fertilizers formulated for low runoff potential.

Chapter | 47  Lawn and Turf: Management and Environmental Issues of Turfgrass Pesticides

Conclusions Turfgrass provides beautiful green areas within our urban and suburban landscapes. Turfgrasses have been maintained by humans to enhance their environment for more than 10 centuries. The complexity and comprehensiveness of the environmental benefits of turfgrass that improves our quality of life are just now being documented through research. A properly planned and maintained turfgrass facility, such as a golf course, offers a diversity of functional benefits to the overall community, in addition to the physical and mental benefits provided by outdoor exercise. Given the pest complex that damages turfgrass, a variety of cultural practices, nutrients, and pesticides may be used to promote turf health. In this chapter, the usages and functional aspects of turf environments were examined to illustrate its uniqueness compared to other plant environments. The uniqueness of the turf environment is particularly amenable to the application of integrated pest management strategies and in the development of best management practices to minimize and mitigate the impact of the pesticides and nutrients used to maintain turfgrasses. Turfgrass has been shown to function efficiently as a means to reduce surface water runoff and can be used in conjunction with vegetative filter strips to provide a barrier against surface and groundwater contamination from pesticides and nutrients used on turf. Furthermore, the increased use of “reduced risk pesticides,” postapplication irrigation, application of pesticides to limited aspects of the turf environment, and reentry intervals following pesticide applications will all reduce the limited human and environmental exposures seen.

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Issues” (J. C. Balogh and W. J. Walker, eds.), p. 221. Lewis, Chelsea, MI. Barbash, J. E., and Resek, E. A. (2000). “Pesticides in Ground Water: Distribution, Trends, and Governing Factors” (R. J. Gilliom, ed.). Lewis, Chelsea, MI. Beard, J. B. (2000). Turfgrass benefits and the golf environmental. In “Fate and Management of Turfgrass Chemicals” (J. M. Clark and M. P. Kenna, eds.), p. 36. American Chemical Society, Washington, DC. Black, K. G., and Fenske, R. A. (1996). Dislodgeability of chlorpyrifos and fluorescent tracer residues on turf: Comparison of wipe and foliar wash sampling techniques. Arch. Environ. Contam. Toxicol. 31, 563. Bowhey, C., McLeod, M. T., and Stephenson, G. R. (1987). Dislodgeable residues of 2,4-D on turf. Proc. Br. Crop Protection Conf. Weeds 8, 799. Brandenburg, R. L., and Villani, M. G. (1998). Behavioral studies of the southern and tawny mole cricket. In “1997 Turfgrass and Environmental Research Summary” (M. P. Kenna and J. T. Snow, eds.), p. 92. U.S. Golf Association, Far Hills, NJ. Branham, B. E., Smitley, D. R., and Miltner, E. D. (1993). Pesticide fate in turf: Studies using model ecosystems. In “Pesticides in Urban Environments: Fate and Significance” (K. D. Racke and A. R. Leslie, eds.), p. 156. American Chemical Society, Washington, DC. Branham, B. E., Miltner, E. D., Rieke, P. E., Zabik, M. J., and Ellis, B. G. (2000). Groundwater contamination potential of pesticides and fertilizers used on golf courses. In “Fate and Management of Turfgrass Chemicals” (J. M. Clark and M. P. Kenna, eds.), p. 78. American Chemical Society, Washington, DC. Brewer, L. W., Driver, C. J., Kendall, R. J., Lacher, T. E., and Galindo, J. C. (1987). Effects of methyl parathion in ducks and duck broods. Environ. Toxicol. Chem. 7, 391. Brewer, L. W., Hummell, R. A., and Kendall, R. J. (1993). Avian responses to organophosphorus pesticides applied to turf. In “Pesticides in Urban Environments: Fate and Significance” (K. D. Racke and A. R. Leslie, eds.), p. 320. American Chemical Society, Washington, DC. Burkart, M. R., and Kolpin, D. W. (1993). Hydrologic and land-use factors associated with herbicides and nitrate in near-surface aquifers. J. Environ. Qual. 22, 646. Carroll, M. J., Hill, R. L., Raturi, S., Herner, A. E., and Pfeil, E. (2000). Dicamba transport in turfgrass thatch and foliage. In “Fate and Management of Turfgrass Chemicals” (J. M. Clark and M. P. Kenna, eds.), p. 228. American Chemical Society, Washington, DC. Cisar, J. L., and Snyder, G. H. (1993). Mobility and persistence of pesticides applied to a USGA-type green: II. Fenamiphos and fonofos. J. Int. Turf Res. Soc. 7, 978. Cisar, J. L., and Snyder, G. H. (1996). Mobility and persistence of pesticides applied to a USGA-type green: III. Organophosphate recovery in clippings, thatch, soil, and percolate. Crop Sci. 36, 1433. Cisar, J. L., and Snyder, G. H. (2000). Mobility and persistence of pesticides applied to a U.S. Golf Association green: Pesticides in percolate, thatch, soil, and clippings and approaches to reduce fenamiphos and fenamiphos metabolite leaching. In “Fate and Management of Turfgrass Chemicals” (J. M. Clark and M. P. Kenna, eds.), p. 106. American Chemical Society, Washington, DC. Clark, J. M., and Kenna, M. P. (eds.), (2000). “Fate and Management of Turfgrass Chemicals,” American Chemical Society, Washington, DC Clark, J. M., Roy, G. R., Doherty, J. J., Curtis, A. S., and Cooper, R. J. (2000). Hazard evaluation and management of volatile and dislodgeable foliar pesticide residues following application to turfgrass. In “Fate and Management of Turfgrass Chemicals”

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(J. M. Clark and M. P. Kenna, eds.), p. 294. American Chemical Society, Washington, DC. Cohen, S. Z., Nickerson, S., Maxey, R., Dupuy, A. Jr., and Senita, J. A. (1990). A ground water monitoring study for pesticides and nitrates associated with golf courses on Cape Cod. Ground Water Monitor. Rev. 10, 160. Cohen, S. Z., Durborow, T. E., and Barnes, N. L. (1993). Groundwater and surface risk assessments for proposed golf courses. In “Pesticides in Urban Environments: Fate and Significance” (K. D. Racke and A. R. Leslie, eds.), p. 214. American Chemical Society, Washington, DC. Cohen, S. Z., Svrjcek, A., Durborow, T., and Barnes, N. L. (1999). Ground water quality: Water quality impacts by golf courses. J. Environ. Qual. 28, 798. Cooper, R. J. (1993). Volatilization as an avenue for pesticide dissipation. In “International Turfgrass Society Research” (�������������������� N. E. Carrow, ed.), Intertec, Overland Park, KS, p. 116. Cooper, R. J., Jenkins, J. J., and Curtis, A. S. (1990). Pendimethalin volatility following application to turfgrass. J. Environ. Qual. 19, 508. Cowell, J. E., Lottman, C. M., and Manning, M. J. (1991). Assessment of lawn care worker exposure to dithiopyr. Arch. Environ. Contam. Toxicol. 21, 195. Cowell, J. E., Adams, S. A., Kunstman, J. L., and Mueth, M. G. (1993). Comparison of foliar dissipation and turf dislodgeable residue sampling techniques. In “Pesticides in Urban Environments: Fate and Significance” (K. D. Racke and A. R. Leslie, eds.), p. 100. American Chemical Society, Washington, DC. Durborow, T. E., et al. (2000). Calibration and validation of runoff and leaching models for turf pesticides, comparison and monitoring results. In “Fate and Management of Turfgrass Chemicals” (J. M. Clark and M. P. Kenna, eds.) American Chemical Society, Washington, DC. Eitzer, B. D., and Chevalier, A. (1999). Landscape care pesticide residues in residential drinking water wells. Bull. Environ. Contam. Toxicol. 62, 420. Felsot, A. S. (1989). Enhanced biodegradation of insecticides in soil: Implications for agroecosystems. Annu. Rev. Entomol. 34, 453–476. Goh, K. S., Edmiston, S., Maddy, K. T., and Margetich, S. (1986). Dissipation of dislodgeable foliar residue for chlorpyrifos and dichlorvos treated lawn: implication for safe reentry. Bull. Environ. Contam. Toxicol. 37, 33. Gold, A. J., Morton, T. G., Sullivan, W. M., and McClory, J. (1988). Leaching of 2,4-D and dicamba from home lawns. J. Water Air Soil Pollut. 37, 121. Golf Course Superintendents Association of America (1998). “1998 Plant Protection and Fertilizer Usage Report.” Golf Course Superintendents Association of America, Lawrence, KS. Grue, C. E., Fleming, W. S., Busby, D. G., and Hill, E. F. (1983). Assessing hazards of organophosphate pesticides to wildlife. Trans. North Am. Wildl. Nat. Res. Conf. 48, 200. Haith, D. A., DiSante, C. J., Roy, G. R., and Clark, J. M. (2000). Modeling approaches for assessment of exposure from volatilized pesticides applied to turf. In “Fate and Management of Turfgrass Chemicals” (J. M. Clark and M. P. Kenna, eds.), p. 255. American Chemical Society, Washington, DC. Harrison, S. A., Watschke, T. L., Mumma, R. O., Jarrett, A. R., and Hamilton, G. W. Jr. (1993). Nutrient and pesticide concentrations in water from chemically treated turfgrass. In “Pesticides in Urban Environments: Fate and Significance” (K. D. Racke and A. R. Leslie, eds.), p. 208. American Chemical Society, Washington, DC.

Hayes’ Handbook of Pesticide Toxicology

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Chapter | 47  Lawn and Turf: Management and Environmental Issues of Turfgrass Pesticides

Murphy, K. C., Cooper, R. J., and Marshall, C. J. (1996b). Volatile and dislodgeable residues following trichlorfon and isazofos application to turfgrass and implications for human exposure. Crop Sci, 36, 1446. Niemczyk, H. D., and Chapman, R. A. (1987). Evidence of enhanced degradation of isofenphos in turfgrass thatch and soil. J. Econ. Entomol. 80, 880. Petrovic, A. M., Young, R. G., Ebel, J. G., and Lisk, D. J. (1993). Conversion of triadimefon fungicide to triadimenol during leaching through soils. J. Chemosphere 26, 1549. Potter, D. A. (1993). Pesticide and fertilizer effects on beneficial invertebrates and consequences for thatch degradation and pest outbreaks in turfgrass. In “Pesticides in Urban Environments: Fate and Significance” (K. D. Racke and A. R. Leslie, eds.), p. 331. American Chemical Society, Washington, DC. Potter, D. A. (1998). Cultural control, risk assessment, and environmentally responsible management of white grubs and cutworms. In “1997 Turfgrass and Environmental Research Summary” (M. P. Kenna and J. T. Snow, eds.), p. 39. U.S. Golf Association, Far Hills, NJ. Racke, K. D. (2000). Pesticides for turfgrass pest management: Uses and environmental issues. In “Fate and Management of Turfgrass Chemicals” (J. M. Clark and M. P. Kenna, eds.), p. 45. American Chemical Society, Washington, DC. Racke, K. D., and Leslie, A. R. (1993). “Pesticides in Urban Environments: Fate and Significance” (M. J. Comstock, ed.). American Chemical Society, Washington, DC. Rainwater, T. R., Leopold, V. A., Hooper, M. J., and Kendall, R. J. (1995). Avian exposure to organophosphorus and carbamate pesticides on a coastal South Carolina golf course. Environ. Toxicol. Chem. 14, 2155. Reinert, J. A. (1982). A review of host resistance in turfgrasses to insects and acarines with emphasis on the southern chinch bug. In “Advances in Turfgrass Entomology” (H. D. Niemczyk and B. G. Joyner, eds.), p. 71. Hammer Graphics, Piqua, OH. Schumann, G. L., Vittum, P. J., Elliot, M. L., and Cobb, P. P. (1997). “IPM Handbook for Golf Courses,” Ann Arbor Press, Chelsea, MI. Schumann, G. L., Clark, J. M., Doherty, J. J., and Clarke, B. B. (2000). Application of DMI fungicides to turfgrass with three delivery systems. In “Fate and Management of Turfgrass Chemicals” (J. M. Clark and M. P. Kenna, eds.), p. 150. American Chemical Society, Washington, DC. Sears, M. K., Bowhey, C., Braun, H., and Stephenson, G. R. (1987). Dislodgeable residues and persistence of diazinon, chlorpyrifos and isofenphos following their application to turfgrass. Pestic. Sci. 20, 223. Shuman, L. M., Smith, A. E., and Bridges, D. C. (2000). Potential movement of nutrients and pesticides following application to golf courses. In “Fate and Management of Turfgrass Chemicals” (J. M. Clark and M. P. Kenna, eds.), p. 78. American Chemical Society, Washington, DC. Smiley, R. W., and Fowler, M. C. (1986). Turfgrass thatch components and decomposition rates in long-term fungicide plots. J. Agron. 78, 633. Smiley, R. W., Fowler, M. C., Kane, R. T., Petrovic, A. M., and White, R. A. (1985). Fungicides on Kentucky bluegrass turf: II. Further observations of effects on thatch and turfgrass growth. J. Agron. 77, 597. Smith, A. E., and Tillotson, W. R. (1993). Potential leaching of herbicides applied to golf courses. In “Pesticides in Urban Environments: Fate and Significance” (K. D. Racke and A. R. Leslie, eds.), p. 168. American Chemical Society, Washington, DC. Smith, C. M., Bottcher, A. B., Campbell, K. L., and Thomas, D. L. (1991). Field testing and comparison of the PRZM and GLEAMS models. Trans. ASAE 34, 838.

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Spencer, W. F., and Cliath, M. M. (1990). Movement of pesticides from soil to atmosphere. In “Long Range Transport of Pesticides” (D. A. Kurtz, ed.), p. 1. Lewis, Chelsea, MI. Spencer, W. F., Farmer, W. J., and Cliath, M. M. (1973). Pesticide volitizaton. Residue Rev. 49, 1. Squillace, P. J., and Thurman, E. M. (1992). Herbicide transport in rivers—Importance of hydrology and geochemistry in nonpoint-source contamination. Environ. Sci. Technol. 26, 538. Starrett, S. K., and Starrett, S. K. (2000). FTURF: Pesticide and nitrogen leaching model. In “Fate and Management of Turfgrass Chemicals” (J. M. Clark and M. P. Kenna, eds.), p. 180. American Chemical Society, Washington, DC. Staten, G. (1946). Contamination of cotton fields by 2,4-D or hormone type weed sprays. J. Am. Soc. Agron. 38, 922. Stone, W. B. (1979). Poisoning of wild birds by organophosphate and carbamate pesticides. N. Y. Fish Game J. 26, 37. Stone, W. B., and Gradoni, P. B. (1985). Recent poisonings of wild birds by diazinon and carbofuran. Northeast Environ. Sci. 4, 30. Taylor, A. W. (1978). Post-application volatilization of pesticides under field conditions. J. Air Pollut. Control 28, 922. Taylor, A. W., Glotfelty, D. E., Glass, B. L., Freeman, H. P., and Edwards, W. M. (1976). Volatilization of dieldrin and heptachlor residues from maize field. J. Agric. Food Chem. 24, 625. Taylor, A. W., Glotfelty, D. E., Turner, B. C., Silver, R. E., and Freeman, H. P. (1977). Volatilization of dieldrin and heptachlor residues from field vegetation. J. Agric. Food Chem. 25, 542. Templeton, S. R., Zilberman, D., and Yoo, S. J. (1998). Pesticides: An economic perspective on outdoor residential pesticide use. Environ. Sci. Technol. News 32, 416. Thompson, D. G., Stephenson, G. R., and Sears, M. K. (1984). Persistence, distribution and dislodgeable residues of 2,4-D following its application to turfgrass. Pestic. Sci. 15, 353. Throssell, C. S., Lyman, G.T., Johnson, M.E., Stacey, G.A., and Brown, C.D. (2009). Golf course environmental profile measures water use, source, cost, quality, and management and conservation strategies. Appl. Turfgrass Sci. doi:10.1094/ATS-2009-0129-01-RS. Thurman, E. M., Goolsby, D. A., Meyer, M. T., Mills, M. S., Pomes, M. L., and Kolpin, D. W. (1992). A reconnaissance study of herbicides and their metabolites in surface water of the midwestern United States using immunoassay and gas chromatography/mass-spectrometry. Environ. Sci. Technol. 26, 2440. Tomimori, S., Nagaya, Y., and Taniyama, T. (1994). Water pollution caused by agricultural chemicals and fertilizers in the drainage from golf courses. Jpn. J. Crop Sci. 63, 442. University of Massachusetts-Extension (1999). “Professional Guide for IPM in Turf for Massachusetts,” University of Massachusetts, Amherst. U.S. Environmental Protection Agency (EPA) (1989). “Exposure Factor Handbook—Appendix 3,” Office of Health and Environmental Assessment. U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (EPA) (1990). “National survey of pesticides in drinking water wells: Phase I report,” EPA 570/99-90-015. Office of Pestices and Toxic Substances, U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (EPA) (1993). “Office of Pesticide Programs Reference Dose Tracking Report,” U.S. EPA, Washington, DC. Watschke, T. L., Mumma, R. O., Linde, D. T., Borger, J. A., and Harrison, S. A. (2000). Surface runoff of selected pesticides applied to turfgrass. In “Fate and Management of Turfgrass Chemicals” (J. M. Clark and M. P. Kenna, eds.). American Chemical Society, Washington, DC.

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Wauchope, R. D., Buttler, T. M., Hornsby, A. G., Augustijn-Beckers, P. W., and Burt, J. P. (1992). The SCS/ARS/CES pesticide properties database for environmental decision-making. Rev. Environ. Contam. Toxicol. 123, 1. Weisskopf, C. P., Seiber, J. N., Maizlish, N., and Schenker, M. (1988). Personnel exposure to diazinon in a supervised pest eradication program. Arch. Environ. Contam. Toxicol. 17, 201. Whitmore, R. W., Kelly, J. E., Reading, P. L., Brandt, E., and Harris, T. (1993). National home and garden pesticide use survey. In “Pesticides in Urban Environments: Fate and Significance” (K. D. Racke and A. R. Leslie, eds.), p. 18. American Chemical Society, Washington, DC. Wilson, J. D., Thurtell, G. W., Kidd, G. E., and Beauchamp, E. G. (1982). Estimation of the rate of gaseous mass transfer from a surface source plot to the atmosphere. Atmos. Environ. 16, 1861.

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Wilson, J. D., Catachpoole, V. R., Denmead, O. T., and Thurtell, G. W. (1983). Verification of a simple micrometeorological method for estimating the rate of gaseous mass transfer from the ground to the atmosphere. Agric. Meteorol. 29, 183. Yeary, R. A., and Leonard, J. A. (1993). Measurement of pesticides in air during application to lawns, trees, and scrubs in urban environments. In “Pesticides in Urban Environments: Fate and Significance” (K. D. Racke and A. R. Leslie, eds.), p. 275. American Chemical Society, Washington, DC. Zinkl, J. G., Rathert, J., and Hudson, P. R. (1978). Diazinon poisoning in wild Canada geese. J. Wildl. Manage. 42, 406. Zweig, G., Leffingwell, J. T., and Popendorf, W. (1985). The relationship between dermal pesticide exposure by fruit harvesters and dislodgeable foliar residues. Environ. Health B 20, 27.

Chapter 48

Pet Care Products Used for Insect Pest Control J. Driver1, J. Ross2, M. Bigelow Dyk3, F. Guerino4 and L. Holden5 1

Risksciences.net, LLC, Manassas, Virginia risksciences.net, LLC, Carmichael, California 3 University of California–Riverside, Riverside, California 4 Intervet/Schering-Plough Animal Health, Roseland, New Jersey 5 Sielken & Associates, Inc., Bryan, Texas 2

48.1  Introduction Pets provide companionship, safety, and service. Many studies have documented the health and life-extending benefits of pet ownership (CDC, 2009). Pets can help owners to decrease blood pressure, cholesterol levels, triglyceride levels, and feelings of loneliness. They can increase opportunities for exercise, outdoor activities, and socialization and are used in a variety of animal-assisted therapy programs (Burch et al., 1995). Given the realization of the enormous value in pets, there has been a substantial increase in the demand by pet owners for various products related to pet care, including life-stage engineered pet foods and nutritional supplements, pet insurance, routine and advanced veterinary care, and veterinary and consumer products for insect pest control. Because the majority of new diseases emerging or reemerging in the United States are zoonotic (meaning they can be transmitted between human beings and animals) the distribution of companion animals and their owners’ utilization of veterinary care can have an impact on pubic health decisions and associated regulatory policy. Of the 1461 microbial and parasitic diseases now recognized in humans, approximately 60% are due to multi-host pathogens characterized by their movement across species lines (Torrey et al., 2005). Over the last three decades, approximately 75% of new emerging human infectious diseases are defined as zoonotic (Taylor

Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

et al., 2001). Our increasing interdependence with animals and their ­ products is an important risk factor for our own health and well-being with regard to infectious diseases. In the case of companion animals, or domestic pets, various external parasites (e.g., fleas, ticks, ear mites, sarcoptic mange mites, demodectic mange mites), internal parasites (e.g., roundworms, hookworms, whipworms, tapeworms, heartworms, coccidian, giardia), and viruses (e.g., West Nile virus) require pest control products, such as insecticides, for prevention and/or treatment. Some of the common infectious agents, such as roundworms and hookworms, are zoonotic. The prevention, control, and/or eradication of pests such as fleas from pet animals and residential settings necessitate the use of products containing insecticides. These products can be in various forms, such as mousses, spot-ons, flea collars, oral tablets, powders, and spray mists. While there are many relatively safe insecticide products available for use on pets, caution still must be observed and products should always be used strictly according to their label directions. Insect growth regulators such as lufenuron, methoprene, and pyriproxyfen can be used in combination or alone with insect (e.g., flea) control products. In the case of fleas, growth regulators can help break the life cycle by inhibiting maturation. Growth regulators have minimal adverse effects by themselves and can be used in combination with adult flea insecticides.

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48.2  Pet ownership and pet care in the united states The American Veterinary Medical Association (AVMA) sponsors periodic surveys characterizing pet ownership and demographics in the United States. These surveys represent the largest, most statistically accurate, and complete surveys available of the pet-owning public and pet population demographics. The survey instrument was a questionnaire distributed by mail to a random sample of U.S. households. This sample was selected to be representative of all U.S. households with respect to geography, pet care product market size, age of household head, and household size and income. Respondents were asked questions about the number and type of animals owned and about the frequency and type of veterinarian visits. Greater detail was requested concerning the household’s last veterinary medical visit, including the products and services provided. The most recent AVMA survey publication (AVMA, 2007) describes the survey data collected in 2006. Questionnaires were mailed to a sample of 80,000 U.S. households from which 47,842 (59.8%) were completed and returned for analysis. Important general findings of this survey were the following: Of all U.S. households, 59.5% owned pets in 2006. The number of U.S. households owning pets grew to 68.7 million in 2006 from 61.1 million in 2001, an increase of 12.4%. l About two-thirds of pet-owning households, or 64.0%, owned more than one pet in 2006, with 21.2% of households owning at least five pets. l In 2006, nearly half of pet owners, or 49.7%, considered their pets to be family members, and nearly half, or 48.2%, considered their pets to be pets/companions. l Overwhelmingly, women across the age spectrum were the primary caregivers of pets. The study found that 74.5% of pet owners with primary responsibility for their pets were female in 2006. l Five of the top-10 pet-owning states in terms of petowning households can be found in the Northwest— Oregon, Washington, Idaho, Montana, and Wyoming. l

In addition, the survey found that households with children continued to rank the highest in pet ownership in 2006, followed by couples and singles. Similarly, pet ownership was highest among those married, followed by those divorced, widowed, or separated, and those never married. The likelihood of owning a pet increased as the size of the household increased. The largest increase in household pet ownership from previous years was found in retired older couples. In general, pet ownership increased as household income increased. This was also true of factors typically correlated with household income: Household members working full time were most likely to own pets, followed by those not employed, those working part-time, and those retired. Those

that owned a home were more likely to own a pet than those that rented. Pet-owning households were more likely to be employed full-time and less likely to be retired. Pet ownership was also associated with type of dwelling and size of community. People living in mobile homes and houses were more likely to own a pet than those living in duplexes, condominiums, or apartments. Pet owners were more likely to live in communities with populations less than 500,000. As community size increased, pet ownership decreased in 2006. As would be expected, the most common pets were dogs and cats. Among pet owners, 30.2% had dogs only, 24.3% had cats only, 1.0% had birds only, and 0.2% had horses only. The remaining pet-owning households had other various combinations of pets in their home. For example, 25.7% of the pet-owning households (or 15.3% of all U.S. households) own both dogs and cats. Tables 48.1 and 48.2 summarize the survey findings regarding the distribution of veterinary services and products for dogs and cats, respectively. As would be expected, physical

Table 48.1  Distribution of Services and Products Purchased During the Most Recent Veterinary Visit by Dog-Owning Households, 1996–2006 1996 (%)

2001 (%)

2006 (%)

Physical exams

61.0

69.4

70.2

Vaccinations

62.8

63.8

64.4

Drugs or medications

29.4

31.3

26.1

Laboratory tests

15.9

18.6

19.9

Flea or tick products

16.7

19.8

18.9

Emergency care

11.5

10.5

11.0

Deworming

8.8

8.2

7.3

Spay/neuter

7.6

6.1

7.3

Grooming/boarding

6.7

7.5

6.6

Dental care

4.4

5.8

5.6

X-rays

3.4

3.8

4.8

Other surgery

4.2

4.3

4.3

Food

3.7

3.7

4.1

Euthanasia

2.9

2.4

4.0

Hospitalization

3.3

2.6

2.4

Microchip/tattoo

a

0.6

1.3

Alternative therapies

a

0.4

0.6

Behavior counseling

0.4

0.3

0.4

a Not available. Adapted from AVMA (2007).

Chapter | 48  Pet Care Products Used for Insect Pest Control

Table 48.2  Distribution of Services and Products Purchased During the Most Recent Veterinary Visit by Cat-Owning Households, 1996–2006 1996 (%)

2001 (%)

2006 (%)

Physical exams

59.6

67.3

71.3

Vaccinations

61.9

70.5

63.7

Drugs or medications

18.8

18.0

17.4

Laboratory tests

12.5

13.5

17.0

Flea or tick products

12.4

15.5

14.0

Spay/neuter

14.0

13.8

14.0

Emergency care

11.8

9.4

11.2

Deworming

7.3

8.0

6.9

Food

5.2

4.3

6.2

Dental care

4.7

3.5

4.7

Other surgery

5.4

4.3

4.7

Euthanasia

3.1

2.6

4.5

X-rays

2.9

2.9

3.9

Hospitalization

4.2

3.6

3.4

Grooming/boarding

3.2

2.7

3.3

Microchip/tattoo

a

0.3

0.7

Behavior counseling

0.5

0.3

0.6

Alternative therapies

a

0.2

0.3

a

Not available. Adapted from AVMA (2007).

exams and vaccinations are the most common services and have essentially the same frequency for both dogs and cats. For the most part, the distribution of services has remained constant or changed only slightly over the past 10 years.

48.3  Pet care products: how they are used to control insect pests Pet care products used to control insects and other pests rely on a variety of active ingredients, including organophosphates, pyrethrins and synthetic pyrethroids, neonicotinoids, phenylpyrazoles, and insect growth regulators such as pyriproxyfen. Available products use several methods to apply active ingredient to the animal and control its delivery to the target pest. The application methods available to pet owners include collars, dusts, dips, shampoos, spot-on treatments, and oral medications. One of the methods that is assumed to have high postapplication human exposure potential is spot-on treatment. The following is an example of typical use instructions for a spot-on product:

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1. Use only on dogs. Do not use on cats or other animals. 2. The dog should be standing or in a comfortable position for easy application. 3. Remove applicator from package. Hold applicator (tube) in upright position away from your face and snap tip open by bending it and folding it back on itself. 4. Part the dog’s hair and place applicator (tube) tip against the skin at the base of the skull. Squeeze applicator firmly, applying entire contents directly to skin. Avoid contact with treated area (application site) until dry. 5. Only one (1) applicator (tube) is needed per treatment. Reapply every 30 days. When application is complete, follow Pesticide Disposal under Storage and Disposal section as described on the product label. Some spot-on products instruct the user to apply the formulation by parting the hair down to the level of the skin; thus, the product formulation is slowly applied to the skin, avoiding superficial application to the animal’s hair. The product formulation can also be applied evenly to one to three or four spots along the dog’s back, typically beginning between the shoulder blades, squeezing the applicator tube until empty (see Figure 48.1). The drops should be spread over enough surface area to avoid runoff and subsequent underdosing. Additionally, whenever possible, dosing should be restricted to areas that would be difficult for the dog to reach and generally inaccessible for grooming purposes (i.e., intrascapular). Once the whole dose has been delivered, the application area should be left untouched by household members, typically for approximately 1 day. Key information from the pet care product labels used in exposure assessments is presented in Table 48.3, using a spoton product as a hypothetical example. Four product sizes are typically available, e.g., for dogs weighing 2–50 kg. Since different amounts may be handled per day depending on breed, gender, and age of the dog, it is important to be aware of the amount used as shown in Table 48.3. It is important to note that dogs are dosed on a weight basis and that dog weight varies widely with breed. However, dosage normalized to surface area is comparable across dogs of different weight ranges (see Table 48.3, max dosage). Additional factors that may be important to consider in pet care product formulations, such as spot-on products, include the surface tension and viscosity of the formulation, which are determined primarily by the solvent. These two physical properties are shown for a hypothetical spoton formulation in comparison with solvents commonly used in agricultural chemical formulations and in dermal absorption studies (see Table 48.4).

48.3.1  Exposure Monitoring Studies As noted previously, a variety of applications (collars, dips, shampoos, spot-on treatments, and oral medications) are

Hayes’ Handbook of Pesticide Toxicology

1080

1

3

1

2

Do not get the product in your dog’s eyes or mouth

Do not get the product in your dog’s eyes or mouth

For dogs weighing 2.5 lbs to 20 lbs, apply the product to one spot, as shown in the diagram above, onto the skin, squeezing the applicator tube until empty.

For dogs weighing 21 to 55 lbs OR for dogs weighing 56 to 95 lbs, apply the product evenly to three spots along the dog’s back, beginning betweent the shoulder blades and continuing in the order shown in the diagram above, squeezing the applicator tube until empty.

4

3

2

Do not get the product in your dog’s eyes or mouth For dogs weighing over 95 lbs, apply the product evenly to four spots on the dog’s back, beginning between the shoulder blades and continuing in the order shown in the diagram above, squeezing the applicator tube until empty.

Figure 48.1  Example diagram of dog spot-on product application procedure (from http://www.summitvetpharm.com/File/3D%20Insert_v7.pdf).

Table 48.3  Hypothetical Dog Spot-On Product Label Information for Exposure Assessment Dog type

Weight of dog (kg) Product size (ml)

Conc. (g/ml)

Active ingredient (AI) in product (g)

Max. dosage (mg AI/cm2)

Very small dog

2–4.5

0.125

0.0563

0.034

Small dog

0.45

4.5–11

1.1

0.125

0.138

0.044

Medium dog

11–22

2.2

0.125

0.275

0.048

Large dog

22–50

5.0

0.125

0.625

0.069

Table 48.4  Physical Properties of Hypothetical SpotOn Formulation Compared with Common Solvents Material

Surface tension Viscosity (mN/m) (mPa s) at (temp) at (temp)

Water

1.0 (20°C)

72.0 (20°C)

Acetone

0.3 (20°C)

23.0 (25°C)

Isopropanol

2.9 (20°C)

20.9 (25°C)

Hypothetical spot-on formulation

8.0 (20°C)

30.0 (25°C)

used. These treatments include an array of pesticide active ingredients, including organophosphates, pyrethroids, neonicotinoids, phenyl and pyrazoles. Because pet owners have extensive contact with their pets, human exposure to these active ingredients is inevitable. A pet as a potential point source for human exposure is a relatively new concept in pesticide exposure assessment.

Some exposure monitoring studies are not focused on the exposure to humans from pet care products per se. Rather, they are concerned with the potential for pets to transfer pesticide residues from the outdoor environment to their owners’ indoor environment. For example, the potential for transfer of diazinon from outdoor turf to indoor residences from dogs was investigated by Morgan et al. (2001). This pilot study measured residues outside the home as well as transferable residues from turf and entryways. They used fur clippings, fur wipes, and paw wipes to measure residues from an indoor/outdoor dog in the home. Other measures of diazinon exposure from the mother, two children, and indoor environment were conducted; however, biomonitoring was not used to estimate exposure in this study. The results showed that the dog was exposed to greater amounts of diazinon than the adult and child occupants and suggested the dog was a good vehicle for uptake, transfer, and translocation of residues into the home. This study was followed by a larger-scale study that also incorporated biomonitoring and transferable residues using gloves (Morgan et al., 2008). The results reported

Chapter | 48  Pet Care Products Used for Insect Pest Control

Amount of Pet Pesticide Product Transferred to Gloves

6000 Total µg pesticide transferred per unit time

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Frontline Plus (Fipronil) Revolution TM (Selamectin)

5000

Advantage (Imidacloprid) Adams Flea and Tick Dip (Chlorpyrifos)

4000

3000

2000

1000

0

Pre

4h

24 h

72 h

1 week 2 weeks 3 weeks 4 weeks 5 weeks Time

Figure 48.2  Potential transferability of pet pest products from dogs indicated by residue on 100% cotton gloves following 5-min petting. Fipronil (Jennings et al., 2002), imidacloprid (Craig et al., 2005), and selamectin (Gupta et al., 2005) data were derived using published data and average glove weight (9.5 g; Gupta, 2008, personal communication). Chlorpyrifos data represent four consecutive treatments with flea control dip (Boone et al., 2001); other products were applied as spot-on treatments (Bigelow Dyk et al., 2009, personal communication).

again showed the dog as a vehicle of transfer of diazinon from the outdoor to indoor environment. Glove residue data showed the diazinon residue transferability from the dog was less transferable after day 1 postapplication. Biomonitoring did not indicate significantly higher levels of diazinon exposure post-treatment and could not be used to link biomarker levels to pet-borne pesticide levels. These studies show the potential for a pet to transfer pesticides that were applied outdoors, but do not indicate the potential for human exposure following applications of pesticides directly to pets. Other studies have been more focused on human exposure directly from pet care applications. Boone et al. (2001) measured transferable chlorpyrifos residues from dog fur to estimate the potential for human exposure from dogs treated with flea dips. This study was the first published of its kind, and the methodology used in this publication was the basis for future transferable pet residue studies. Boone et al. (2001) used 100% cotton gloves as a wipe to collect transferable residues from dogs after they were treated with flea dip containing chlorpyrifos. Briefly, investigators wearing cotton gloves petted dogs for 5 min. Gloves were extracted as a measure of transferable residue. This study showed the potential for chlorpyrifos residues to be transferred for up to 21 days post use of flea dips (Figure 48.2). The interpretation and extrapolation of exposure monitoring results are dependent on the particular measurement method(s) used. The use of cotton gloves as a dosimeter for transferable residues has become commonplace in pet

pesticide exposure assessment. The procedure has not been validated as a surrogate measure of human exposure; that is, there are no data linking human exposure to transferable residues. Figures 48.3 and 48.4 show the use of a mannequin hand (Northwest Mannequin, Seattle, WA; part number Hand M2) used to conduct simulated petting on a treated dog. The mannequin hand is covered with a nitrile glove (see Figure 48.3) and then two layers of cotton gloves (see Figure 48.4). In this manner, residues transferred to the cotton glove can be measured as a function of the number of petting contacts and as a function of time postapplication to the pet. After Boone et al. (2001), six additional studies have been published using gloves as a dosimeter of exposure (Boone et al., 2002; Chambers et al., 2007; Craig et al., 2005; Davis et al., 2008; Gupta et al., 2005; Jennings et al., 2002). All of these studies show the potential transferability of pesticide residues, but they fail to directly link transferable pesticide residues with human absorption. Two studies on flea collars containing tetrachlorvinphos and chlorpyrifos were published with biomonitoring data (Chambers et al., 2007; Davis et al., 2008). These studies also included the use of t-shirts on children as an indicator of transferability from the pet to children. Residues were observed on the t-shirts (ng/g chlorpyrifos; g/g tetrachlorvinphos). Post-treatment residues were significantly higher than residues pretreatment, but there was no significant difference observed over time. Biomarker levels were not significantly higher than background levels attributed to dietary sources of organophosphates. These studies did not

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Figure 48.3  Mannequin hand used with cotton glove dosimeters to measure transferable residues from treated pets.

Hayes’ Handbook of Pesticide Toxicology

that would need to be documented in a pet care biomonitoring study. The preceding studies have successfully investigated the potential for transfer of pet product active ingredients from pets to pet owners . However, to date, these studies do not give a complete picture of the pet owners’ actual exposure from these products. Currently, a pilot study sampling pet owners who use Frontline is being conducted. This study will include fipronil residues on pet hair, glove dosimeters, and sock dosimeters to sample indoor surfaces, as well as concurrent urine biomonitoring to give a more comprehensive understanding of the fate of spot-on products and their potential for human exposure (Bigelow Dyk, 2009).

48.3.2  Pet Care Product Applicator and Postapplicator Exposure Assessment Methods

Figure 48.4  Demonstration of mannequin hand during petting simulation study.

show a significant increase in biomarker excretion after application of flea collars to dogs. Recently, spot-on flea and tick treatments have become more available to pet owners for protection against endoand ectoparasites. As noted previously, this method is of special interest to regulators because of a potential for human exposure for several weeks after treatment. Three well-known spot-on treatments (Frontline, Revolution, and Advantage) have been investigated for potential human exposure (Craig et al., 2005; Gupta et al., 2005; Jennings et al., 2002). These studies again demonstrated the transferability of these insecticides using cotton gloves (Figure 48.2). The research has not included human biomonitoring to link pet owner or occupational exposure. Frontline and Revolution would be well suited to biomonitoring studies, since pet owners are not likely exposed to fipronil and selamectin from other sources. In contrast, imidicloprid, the active ingredient in Advantage, is commonly used for purposes of residential lawn and ornamental treatment, and thus would have potential confounding sources of exposure

The U.S. Environmental Protection Agency (EPA) addressed pet care exposures initially as part of the 1996 Food Quality Protection Act (FQPA; http://www.epa.gov/ opp00001/regulating/laws/fqpa/) as part of the issuance of the first version of its Standard Operating Procedures for Residential Exposure Assessment in 1997 (U.S. EPA, 1997). Because additional studies have been conducted since 1997, it is important to make incremental changes in the regulatory approaches used to reflect new data as they become available. The products used to control pests on animals are varied, and the EPA is interested in the use of any of these that result in exposure to adults during application and to adults and children postapplication. It is assumed that small children (e.g., toddlers) potentially have the highest postapplication exposures associated with the use of these types of products compared with adults or older children because of the time spent with animals and the nature of the contact they may have with animals. For example, small children might be more prone to put their hands in their mouth after petting a dog without washing their hands. The EPA has recently developed screening-level standard operating procedures (SOPs) to address consumer (residential) applicator and postapplication exposure scen­ arios, including those associated with different types of pet care products (U.S. EPA, 2009). The SOPs involve a series of assumptions and exposure factors that serve as the basis for estimating potential dermal, inhalation, and ingestion exposures during and following (post) application of a specified product type and formulation. In the case of adult applicator exposure estimation, the following assumptions and factors are included: 1. Daily exposure (and absorbed dose) is based on the amount of active ingredient handled on the day of treatment (i.e., a single pet treatment). The maximum labelspecified application rate is assumed. The maximum

Chapter | 48  Pet Care Products Used for Insect Pest Control

rate is for a single pet and is determined using a representative (average) animal, e.g., a medium-sized dog or cat weighing 30 and 10 lbs, respectively. Estimated risks are based on an assumed homogeneous distribution of residues across the entire surface area1 of the animal. 2. In the case of homeowner applicator exposures during dipping, dusting, or shampooing pets, 10% of the active ingredient applied is assumed to be the amount of exposure. This is assumed to be an upper-percentile estimate. 3. In the case of flea collars, 1% of the active ingredient applied is assumed to be available for dermal and inhal­ ation exposure during pet handling. 4. In the case of spray formulations, it is assumed that the person applying the product uses half of a full can of spray. 5. Dermal absorption (usually expressed as a percentage during the time period of interest, e.g., per daily dermal exposure event) is considered in the exposure assessment if a systemic toxicology endpoint is being considered. 6. It is assumed that applicator contact with a treated pet during application of a spot-on product is negli­gible, given that these products are designed to be selfcontained and are applied directly from the tube to the pet with the tip of the applicator tube, which is used to part the pet’s hair during application. 7. Adults are assumed to weigh 70 kg (mean value for adult males and females); a mean value of 60 kg for females ages 13 to 54 years is used when the selected toxicological endpoint used in risk assessment is from a reproductive or developmental toxicity study. The algorithms recommended by the EPA (U.S. EPA, 2009) for assessing potential dermal and inhalation ex­posures from residential application of pet care products are as follows. For spray formulations:



daily dermal or inhalation exposure (mg / kg / day ) ( UE  AR )  BW

where UE  unit exposure (mg/lb active ingredient or AI), AR  application rate or amount applied to the animal (lbs AI/treatment), and BW  body weight (adults; kg). For other formulations:



1

daily dermal or inhalation exposure (mg / kg / day) (AR  F )  BW

For example, in the case of a 30-lb dog, the equation used to estimate surface area (cm2)  12.3  [(body weight, lbs  454 mg/lb)0.65]  5986 cm2 (U.S. EPA, 1993).

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where AR  application rate or amount applied to the animal (mg AI/treatment), F  fraction of AI applied, which is assumed to be available for dermal or inhalation exposure, and BW  body weight (adults; kg). Although the EPA acknowledges that 10% transfer of product to the applicator is the “upper bound,” the measurements that have been made in the case of spot-on and other formulations used by applicators suggest that that actual exposures are associated with less than 1% of the total amount used (U.S. EPA, 1998). In the case of potential postapplication exposure assessment for adults and children, the EPA (U.S. EPA, 2009) uses a series of assumptions and exposure factors that serve as the basis for derivation of dermal and incidental oral (handto-mouth) exposure estimates from all formulation types. The general assumptions and factor used are as follows: 1. Same as #1 for applicator exposure. 2. On the day of application, it may be assumed that 20% of the maximum application rate is available on the pet’s body and transferable to an adult or child (toddler) as a dislodgeable residue. This value is based on professional judgment and the experience of EPA scientific staff involved in reviewing pesticide registrant-submitted data. It is characterized as an upper-percentile value (U. S. EPA, 2009). 3. If chemical- and use-specific data have been submitted to the EPA (e.g., a transferable residue study measuring residue dislodgeability from a treated pet to a surrogate mannequin hand with cotton gloves during petting simulation), these data will be used rather than default values/ assumptions. Measured values from studies collected on the day of application (“day 0”) are typically used to estimate the percent transfer (pet to hands/body areas). 4. Postapplication activities must be assessed on the same day that the pesticide is applied because it is assumed that individuals could handle/touch their pets immediately after application. For subsequent days post­ application, it may be assumed that residues do or do not dissipate, depending on the product type, e.g., the continuous use of flea collars versus spot treatment at 30-day intervals. 5. It is assumed that one pet is contacted per day. 6. Adults (18 years and older) are assumed to weigh an average of 70 kg (60 kg for 13- to 54-year-old females); a 3-year-old toddler is assumed to weigh 15 kg (average weight for 1- to 6-year-old children). 7. Dermal absorption (usually expressed as a percentage during the time period of interest, e.g., per daily dermal exposure event) is considered in the exposure assessment if a systemic toxicology endpoint is being considered. 8. The EPA assumes that the body surface area of adult and child daily pet “hugging” events is 5625 and 1875 cm2, respectively.

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9. In the case of the hand-to-mouth exposure pathway for children, it is assumed that the saliva removal efficiency of residues on the hand is 50%; that is, every time the hand goes into the child’s mouth, approximately one-half of the residues on the hand are removed. 10. The hand-to-mouth exposure pathway for children is assumed to result from an “equilibrium-based” hand loading. This assessment method is based on the assumption that children may contact the pet repeatedly with their hands and thereby reach a maximal loading on the skin. This is analogous to the dermal exposure pathway assessment approach. Thus, one equilibrium-based hand-to-mouth event is assumed to occur per day. The algorithms used for residential postapplication dermal and incidental oral ingestion (hand-to-mouth) pet exposure scenarios are as follows. For dermal exposure estimation associated with a touching/hugging a treated pet (e.g., dog):



daily dermal exposure (mg / kg / day ) [((AR  F ) / SA pet )  SA hug ]  BW

where AR  application rate or amount applied to the animal (mg AI/treatment), F  fraction of the application rate available as transferable residue (0.2), or fraction determined from a chemical-specific study, SApet  surface area of a treated dog (for a 26-lb dog  5986 cm2), SAhug  surface area of child dermal area (hug) (5625 and 1875 cm2 for adults and children, respectively), and BW  body weight (kg). For incidental oral (hand-to-mouth) exposures attributable to a child touching a treated pet (e.g., dog):



daily oral exposure (mg / kg / day ) [((AR  F ) / SA pet )  SR  SA hands  FR]]  BW

where AR  application rate or amount applied to the animal (mg AI/treatment), F  fraction of the application rate available as transferable residue (0.2), or fraction determined from a chemical-specific study, SApet  surface area of a treated dog (for a 26-lb dog  5986 cm2), SR  saliva removal efficiency (50%), SAhand  surface area of child’s hand involved in mouthing (20 cm2 for adults and children), FR  hand-to-mouth frequency (one event per day, based on equilibrium hand loading of residues from the treated pet), and BW  body weight (kg). In addition to the screening-level deterministic exposure assessment methods described previously, probabil­ istic exposure assessment methods are also important to consider if underlying data sources supporting exposure

assessment input variable distributions are available (U.S. EPA, 2009). In probabilistic analyses, the assessor must establish, characterize, and validate the input distributions used in place of those assumed in screening-level deterministic assessments.

48.4  Case study: illustration of EPA methodology To illustrate deterministic and probabilistic exposure assessment differences for pet care pesticides, a hypothetic­al spoton product for dogs was evaluated using the previously described screening-level algorithms and assumed input variable values, and then compared with a probabilistic assessment. The probabilistic assessment is based on input distributions for transferable residues as would be available from studies conducted for this purpose and distributions of other common exposure factors (e.g., body weight). In contrast to screening-level (deterministic) residential ex­posure and risk assessments often conducted by regulatory agencies, probabilistic assessments can provide insights into the range of potential exposures and an indication of the most sensitive input variables. This example analysis is for potential short-term (day of product application) exposures for the presumed highest exposed subgroup, that is, young children, who interact with the pet dog postapplication of the spot-on product. It is important to note that seasonal, intermediate-term exposures2 to children may also be important to characterize. The case study exposure assessment was conducted using the software program Microsoft Excel in conjunction with Crystal Ball (version 7.2.1). The assessment is based on the use of a combination of guidance from EPA (U.S. EPA, 2009), publicly available data, and hypothetical chemicalspecific data, that is, chemical-specific pet fur transferable residue measurements. Thus, input variable values including distributional representations, when appropriate data were available to specify them, were based on chemicalspecific data, or in the case of general exposure factors, the EPA’s Residential Standard Operating Procedures (SOPs) (U.S. EPA, 1997, 1999a, 2009), including EPA Policy 12recommended scenario-specific defaults (U.S. EPA, 2001), the EPA organophosphate (U.S. EPA, 2002) and N-methyl carbamate cumulative residential risk assessment (U.S. EPA, 2007b) input values, the International Life Sciences Institute, Health and Environmental Sciences Institute’s 2

The EPA and other regulatory agencies use time-averaged exposure metrics for comparison with toxicological endpoints resultant from repeat dosing studies. Examples include lifetime average daily exposure (LADE) or lifetime average daily dose (LADD) estimates for derivation of lifetime excess cancer risk, and subchronic and chronic average daily exposure (or absorbed dose) estimates used to derive intermediate-term and chronic, non-cancer margins of exposure (MOEs). For more discussion, see the EPA’s Guidelines for Exposure Assessment (http://cfpub.epa.gov/ncea/ cfm/recordisplay.cfm?deid  15263).

Chapter | 48  Pet Care Products Used for Insect Pest Control

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Table 48.5  Input Variable Values Used in Probabilistic Pet Care Postapplication Child “Day of Application” Exposure Analysis: Hypothetical Spot-On Product for Dogs—Equilibrium Contact Simulation Input variable

Value(s)

Units

Notes

Amount of AI applied

275

mg A.I./treatment

Product label (for spot-on treatment) (275 mg for 26-lb (11.8-kg) dog)

Fraction of application rate available as transferable residue (AR  F): day of application

0.005, 0.02, 0.2

Unitless

2-h to 24-h pet fur transferability: triangular distribution representing minimum, most likely, and maximum values (i.e., 0.005, 0.02, 0.2) expressed as fraction transferred; maximum value assumed is default 20%

Surface area of treated dog

5987

cm2

5987 cm2 (11.8-kg dog); 10,190 cm2 (22.7-kg dog); EPA SOP (U.S. EPA, 1999a, 2001, 2008, 2009)

Surface area involved in child dermal contact (hug)

938 to 1875

cm2

Uniform distribution; 1875 cm2 from U.S. EPA (1999a, 2009) as an upper bound, and half this default assumed as lower bound

Surface area of child’s hand involved in mouthing behavior

1, 7, 20

cm2

Triangular distribution (min, most likely, max) based on ILSI Expert Committee (ILSI, 2004), EPA OP Cumulative (U.S. EPA, 2002c), and U.S. EPA, 2008

Saliva removal efficiency (extraction factor)

0.5

Unitless

Single value (U.S. EPA, 2009)

Frequency of hand-to-mouth events

1

day1

Single value; one event per day based on equilibrium transfer method

Exposure duration (hand-to-mouth)

1

days

Single value; one event per day based on equilibrium transfer method (U.S. EPA, 2009)

Child body weight

18.9; 1.22

kg

EPA OP Cumulative (U.S. EPA, 2008) (lognormal, with geomean.  18.9 kg; geo std dev  1.22 kg)

Product use events per month

1

month1

One use per month; product label (for spot-on treatment)

Exposure Factors Database (ILSI, 2004), the EPA’s Exposure Factors Handbook (U.S. EPA, 1999b), EPA’s Child-Specific Exposure Factors Handbook (U.S. EPA, 2000), and a recent pet care probabilistic assessment conducted by the EPA (U.S. EPA, 2008). Two alternative probabilistic simulations were conducted to estimate potential dermal and incidental oral (hand-to-mouth) postapplication exposures to children, assuming that they interact with a treated dog on the day of treatment. The first simulation was based on the equilibrium concept and involves a single equilibrium pet contact per day (on hands and body areas). The single equilibrium

c­ ontact actually reflects the cumulative loading (residue transfer) from repeat contacts that would result in equilibrium loading and associated dermal and incidental oral exposures. The input variable values used in the first simulation are presented in Table 48.5. In comparison, Table 48.6 presents an alternative simulation representing dermal and incidental oral (hand-to-mouth) exposures associated with repeat contacts on the day of product application. The results of the probabilistic simulations, in comparison to the screening-level deterministic estimates of dermal and oral daily exposures, are presented in Table 48.7 (equilibrium

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Table 48.6  Input Variable Values Used in Probabilistic Pet Care Postapplication Child Exposure Analysis: Hypothetical Case Study for Dog Spot-On Product—Alternative Simulation (Repeat Contact, Subequilibrium) Input variable

Value(s)

Units

Notes

Amount of AI applied

275

mg A.I./treatment

Product label (for spot-on treatment) (275 mg for 26-lb (11.8-kg) dog)

Fraction of application rate available as transferable residue (F-TR): day of application

0.005, 0.02, 0.1

Unitless

2-h to 24-h results: triangular distribution representing minimum, most likely, and maximum values (i.e., 0.005, 0.02, 0.2) expressed as fraction transferred; maximum value based on realistic experimental max value 0.1 (10%) transfer

Surface area of treated dog

5987

cm2

5987 cm2 (11.8-kg dog); EPA SOP (U.S. EPA, 1999a, 2001, 2008, 2009)

Surface area involved in child dermal contact (hug)

938 to 1875

cm2

Uniform distribution; 1875 cm2 from EPA SOPs (U.S. EPA, 1999a) as an upper bound, and half this default assumed as lower bound

Surface area of child’s hand involved in mouthing behavior

1, 7, 20

cm2

Triangular distribution based on ILSI Expert Committee (ILSI, 2004), EPA OP Cumulative (U.S. EPA, 2002c), and U.S. EPA (2008)

Saliva removal efficiency (extraction factor)

0.2 to 0.5

Unitless

Uniform distribution; based on chemicalspecific data from sources such as human in vitro dermal wash off at 24 h

Frequency of hand-to-mouth events

1, 7, 13

h1

Triangular (1, 7, 13) assumes that 50% of total hourly hand-to-mouth events are petrelated; total hourly frequency is represented as a triangular distribution of 1, 13, 26 events/h (ILSI, 2004; Tulve et al., 2002; U.S. EPA 2008, 2002c; Xue et al., 2007)

Exposure duration (hand-to-mouth)

0.03, 0.11, 1.03

day1

Triangular based on EPA OP Cumulative best estimate value of 0.108 h, with min and max of 0.03 and 1.03 h (U.S. EPA, 2008)

Child body weight

18.9, 1.22

kg

EPA OP Cumulative (lognormal: GM  18.9 kg, GSD  1.22 kg)

Product use events per month

1

month1

One use per month; product label (for spoton treatment)

contact) and Table 48.8 (repeat, subequilibrium contact). The simulation results suggest that screening-level deterministic estimates are, as intended, upper-bound estimates. The percentile estimates are based on a realistic input distribution for either equilibrium-based (Table 48.7) or repeat contact (Table 48.8) transferable residue values (percent transfer) during the first 24 h postapplication. Transferable residues are the most sensitive input variable for both the dermal and oral exposure estimates; the second most sensitive variables for the dermal and oral routes, respectively, are the surface areas of the child exposed dermally and the surface area of the hand involved in mouthing. Several factors combine to render pet care assessments conservative; that is, there is a tendency to over­estimate exposure. Enumerated below are the variables that make this assessment an overestimate of exposure and an

approximate numerical indicator of that overestimate is provided in Table 48.9.

48.4.1  Properties of Formulation The dermal absorption guidelines (OECD, 2002; U.S. EPA Health Effects Test Guidelines for dermal penetration studies OPPTS 870.7600; Zendzian, 1994) state that the test material is to remain in place 8 h or more before washoff. However, pet care product labels often instruct the user to wash their hands after use, and postapplication transfer to skin would likely be followed by hand washing, which typically occurs at least daily. Hand washing would be expected to remove, to some extent, formulation and associated active ingredient residues on the skin.

Chapter | 48  Pet Care Products Used for Insect Pest Control

Table 48.7  Percentile Estimates of Child Dermal and Incidental Daily Postapplication Exposure in Comparison to Screening-Level Deterministic Estimates: Hypothetical Spot-On Product for Dogs— Equilibrium Contact Simulation Percentiles

Child incidental oral Child dermal exposure (day 0) exposure (day 0) (mg/kg/day) (mg/kg/day)

0.1%

0.019

0.00003

1%

0.029

0.0001

5%

0.050

0.0001

10%

0.069

0.0002

50%

0.220

0.0007

75%

0.360

0.0012

90%

0.507

0.0018

95%

0.599

0.0022

99%

0.777

0.0032

99.9%

0.963

0.0044

Tier 1 screeninglevel deterministic estimates

1.15

0.006

48.4.2  Postapplication Dermal and Incidental Ingestion Exposure Variables The dermal and oral exposures/doses estimated for children are likely to be overestimates. This is due to the likely conservative bias associated with multiple input variable values, including the dermal and hand transferable residues, skin surface area assumed to be exposed, the dermal absorption distribution used, hand-to-mouth frequency, and others. Conservative biases in exposure factors have been discussed by various expert committees (ILSI, 2004; U.S. EPA, 1997, 2001, 2002b).

48.4.3  Activity Patterns It was conservatively assumed that children would interact and substantially contact their family’s pet dog every day. This does not account for time spent away (U.S. EPA, 1999b, 2000) from home during the season of spot-on product use.

48.4.4  Research Data and Assessment Methodology Needs for Pet Care Pesticide Products Key input variables in the EPA’s predictive algorithms for both potential dermal (adults and children) and incidental

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Table 48.8  Percentile Estimates of Child Dermal and Incidental Daily Postapplication Exposure in Comparison to Screening-Level Deterministic Estimates: Hypothetical Spot-On Product for Dogs— Repeat Contact, Subequilibrium Simulation Percentiles

Child dermal exposure (day 0) (mg/kg/day)

Child incidental oral exposure (day 0) (mg/kg/ day)

0.1%

0.017

0.00001

1%

0.026

0.00003

5%

0.040

0.0001

10%

0.052

0.0001

50%

0.126

0.0006

75%

0.194

0.0012

90%

0.265

0.0021

95%

0.311

0.0030

99%

0.409

0.0053

99.9%

0.538

0.0105

Tier 1 screeninglevel deterministic estimates

1.15

0.006

ingestion (assumed to occur in children, e.g., ages 2 to 6 years) exposure estimation include: 1. the amount applied, deposited, and translocated (residue migration) on a pet’s surface area as a function of time postapplication; 2. the fraction of residues applied (deposited) on a pet (dog or cat) that might transfer to dermal contact areas, e.g., body areas such as arms, torso, and legs during pet contact, and to hands during petting; 3. the surface area of hands and other body areas contacting treated animals and the frequency and duration of contact; and 4. the frequency and duration of hand-to-mouth events in the case of children assumed to be engaging in this behavior. While the EPA’s algorithms include assumptions for some of these input variable values for purposes of deterministic, screening-level calculations, distributions for some of these variables can also be developed and used in probabilistic analyses. One of the most sensitive input variables is the fraction of the applied active ingredient that becomes a transferable residue. Various transferable residue measurement methods have been explored by investigators to approximate equilibrium percent transfer. As noted previously, the EPA’s current dermal and oral ­ estimation

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Table 48.9  Exposure Factors Producing Potential Estimation Bias Exposure factor

Default used in case study assessment

Realistic factor

Overestimate bias (approx.)

Reference

Wash-off from hands

None

50%

2

U.S. EPA (1999b)

2

2

Finger surface area put in mouth

20 cm

7 cm

3

ILSI (2004)

Surface area contacting dog

1875 cm2

Hands, arms, and face only

2

U.S. EPA (1999b)

Contact over a 90-day period

90 days

Time spent inside/outside residence minus time away from home (e.g., in transit, visiting, at school)

1.2

Consumer product use and time activity pattern data (U.S. EPA, 1999a)

Clothing penetration

50%

10%

5

Driver et al. (2007)

Benchmark Contact Level (e.g., 15 Petting Simulations)

Exceeds dog’s tolerance – cannot achieve experimentally

Contact Level at Equilibrium% Transfer (may or may not achieve this experimentally)

Total % Transfer

1 day

4 days

14 days

28 days

Amount. of Pet Contact (No. of Petting Simulations, e.g., 2, 4, 8, 16, 32) Figure 48.5  Illustration of insecticide transferability as function of pet contact (i.e., simulations, wherein one simulation represents three to five petting strokes of a treated animal at a given time postapplication).

algorithms are based on an assumed equilibrium transferability metric (percent transfer) – see Figures 48.5 and 48.6. Figure 48.5 illustrates the general nature of transfer from animals to owners. It indicates that total cumulative transfer increases with amount of pet contact at a particular time. At some point, increasing the amount of pet contact does little to promote additional transfer. This results in a “theoretical maximum” or “equilibrium” level of residue transfer from the treated pet to a person contacting the pet. Figure 48.5 illustrates challenges with experimental attempts to simulate different degrees of pet contact. For example, the experimental animal may not be able to tolerate the

degree of contact (a high number of repeat petting simulations) required to reach equilibrium levels. Thus, a smaller, experimentally feasible number of petting simulations is necessary. However, this is not an issue if the rate of transfer is all that is needed, that is, the nonequilibrium method of exposure estimation. Finally, it is important to recognize that actual human–pet contact levels may typically be below the equilibrium levels. Figure 48.6 illustrates that for any particular level of contact, if measured over time, the percent transfer typically declines exponentially. The rate of transfer decrease for each benchmark (i.e., the slope of the line) is about the same. Thus, if the purpose is to ­measure

Chapter | 48  Pet Care Products Used for Insect Pest Control

Equilibrium Level of Pet Contact

Log Total % Transfer

C

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B

Rate of Decrease in Transfer with Time can be Measured at any Level of Pet Contact

Pre-Equilibrium Levels of Pet Contact A

Days After Treatment Figure 48.6  Illustration of the rate of decrease in pet fur residue transfer as a function of days post-treatment at low (A), medium (B), and high (C; at equilibrium) contact levels.

the rate of decrease, any benchmark level of contact (low, medium, or high, i.e., at equilibrium) can be used.

Conclusion In conclusion, key areas of research for pet care pesticide products include the following: 1. pet care product use/usage data collected in conjunction with human–pet interaction videography and/or diary surveys; 2. comparative evaluation of transferable residue methods (e.g., human hands, gloved human hands, gloved mannequin hands) and associated modeling of the data as a function of contact level, contact duration, and time post-treatment; and 3. evaluation of predictive algorithms with comparisons to the results of human exposure and biomonitoring studies. Potential consumer, residential pet care product-related exposure can be estimated using varying levels of refinement. The ideal methods for estimating exposure involve chemicalspecific passive dosimetry that measures dermal and inhal­ ation exposure or biomonitoring (Ross et al., 2008). Pet care products provide valuable companion animal disease prevention and human public health intervention. Understanding animal safety and potential human health risks associated with these products is a very important ongoing process that will continue to be used to balance risks and benefits.

References American Veterinary Medical Association (AVMA), 2007. “U.S. Pet Ownership and Demographics Sourcebook,” American Veterinary Medical Association. Schaumburg, Illinois. ISBN-10: 1882691172; ISBN-13: 978-1882691173.

Bigelow Dyk, M. M. (2009). “Determinants of Human Exposure to Fipronil Following use as a Topical Flea and Tick Treatment of Companion Animals,” University of California, Riverside, CA. Boone, J., Tyler, J., and Chambers, J. (2001). Transferable residues from dog fur and plasma cholinesterase inhibition in dogs treated with a flea control dip containing chlorpyrifos. Environ. Health Perspect. 109, 1109–1114. Boone, J. S., Tyler, J., and Chambers, J. E. (2002). Exposure of children and adults to transferable residues of chlorpyrifos from dogs treated with flea control collars. Toxicol. Sci. 66(suppl), 346. Burch, M. R., Bustad, L. K., Duncan, S. L. et al. (1995). The role of pets in therapeutic programmes. In “The Waltham Book of HumanAnimal Interaction: Benefits and Responsibilities of Pet Ownership” (I. Robinson, ed.), pp. 55–69. Elsevier Science Inc, Tarrytown, NY. CDC (Centers for Disease Control and Prevention). 2009. “Health Benefits of Pets,” http://www.cdc.gov/HEALTHYPETS/health_benefits.htm Accessed June 2009 Chambers, E., Boone, J., Davis, M., Moran, J., and Tyler, J. (2007). Assessing transferable residues from intermittent exposure to flea control collars containing the organo-phosphate insecticide chlorpyrifos. J. Expo. Sci. Environ. Epidemiol. 17(7), 656–666. Craig, M., Gupta, R., Candery, T., and Britton, W. (2005). Human exposure to imidacloprid from dogs treated with Advantage. Toxicol. Mech. Methods 15, 278–291. Davis, K., Boone, J., Moran, J., Tyler, J., and Chambers, J. (2008). Assessing intermittent pesticide exposure from flea control collars containing the organophosphorus insecticide tetrachlorvinphos. J. Expo. Sci. Environ. Epidemiol. 18, 564–570. Driver, J., Ross, J., Mihlan, G., Lunchick, C., and Landenberger, B. (2007). Derivation of single layer clothing penetration factors from the pesticide handlers exposure database. Regul. Toxicol. Pharmacol. 49, 125–137. Gupta, R., Masthay, M., Canerdy, T., Acosta, T., Provost, R., Britton, D., Atieh, B., and Keller, R. (2005). Human exposure to selamectin from dogs treated with RevolutionTM: methodological consideration for selamectin isolation. Toxicol. Mech. Methods 15(4), 317–321. ILSI. (2004). “HESI Residential Exposure Factors Database Users Guide,” ILSI, Health and Environmental Sciences Institute, Washington, DC. http://hesi.ilsi.org/index.cfm?pubentityid  47.

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Jennings, K., Canerdy, T., Keller, R., Atieh, B., Doss, R., and Gupta, R. (2002). Human exposure to fipronil from dogs treated with Frontline. Vet. Hum. Toxicol. 44(5), 301–303. Morgan, M., Stout, D., and Wilson, N. (2001). Feasibility of the potential for human exposures to pet-borne diazinon residues following lawn applications. Bull. Environ. Contam. Toxicol. 66, 295–300. Morgan, M., Stout, D., Jones., P., and Barr, D. (2008). An observational study of the potential for human exposures to pet-borne diazinon residues following lawn applications. Environ. Res. 107, 336–342. Ross, J., Chester, G., Driver, J., Lunchick, C., Holden, L., Rosenheck, L., and Barnekow, D. (2008). Comparative Evaluation of Absorbed Dose Estimates Derived from Passive Dosimetry Measurements with Those Derived From Biological Monitoring: Validation of Exposure Monitoring Methodologies. J. Expo. Sci. Environ. Epidemiol. 18, 211–230. Taylor, L. H., Latham, S. M., and Woolhouse, M. E. (2001). Risk factors for human disease emergence. Philos. Trans. R Soc. Lond. B Biol. Sci. 356, 983–989. Torrey, E. F., and Yolken, R. H. (2005). “Beasts of the Earth,” Rutgers University Press, New Brunswick, NJ. Tulve, N. S., Suggs, J. C., McCurdy, T., Cohen Hubal, E. A., and Moya, J. (2002). Frequency of mouthing behavior in young children. J. Expo. Anal. Environ. Epidemiol. 12, 259–264. U.S. EPA (U.S. Environmental Protection Agency). (1993). Wildlife Exposure Factors Handbook. Office of Health and Environmental Assessment, Office of Research and Development, U.S. Environmental Protection Agency, Washington, D.C. 20460. EPA/600/R-93/187. U.S. EPA (U.S. Environmental Protection Agency). (1997). Standard Operating Procedures for Residential Exposure Assessments. 6/12/97. U.S. EPA (U.S. Environmental Protection Agency). 1998. PP#7F04832. ID#000264-00554. ID#065331-00003. ID#000264-LTT. Fipronil for use on Rice (Regent®, Icon®) and Pets (Frontline®). HED Risk Assessment. Chemical 129121. Barcodes D242090, D245656, D245627, & D241676. Cases 288765, 031271, 060305, & 061662. Submissions S535722, S541670, S541551, S534929. Memorandum from George F. Kramer et al. to Susan Lewis dated 22 May, 1998. U.S. EPA (U.S. Environmental Protection Agency). (1999a). Overview of Issues Related to the Standard Operating Procedures for Residential Exposure Assessment-Presentation to the FIFRA Science

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Advisory Board, September 21, 1999. Office of Pesticide Programs, Washington, D.C. U.S. EPA (U.S. Environmental Protection Agency). (1999b.) (February) Exposure Factors Handbook. EPA/600/C-99/001 (CD Rom). National Center for Environmental Assessment, Cincinnati, OH. U.S. EPA (U.S. Environmental Protection Agency). (2000). ChildSpecific Exposure Factors Handbook. NCEA-W-0853. June 2000. External Review Draft. U.S. EPA (U.S. Environmental Protection Agency). (2001). Science Advisory Council for Exposure Policy Number 12: Recommended Revisions to the Standard Operating Procedures (SOPs) for Residential Exposure Assessments. Revised February 22, 2001. U.S. Environmental Protection Agency, Office of Pesticide Programs, Washington, D.C. U.S. EPA (U.S. Environmental Protection Agency). (2002). Revised Organophosphate Pesticides (OP) Cumulative Risk Assessment (see Residential Chapter). http://www.epa.gov/pesticides/cumulative/rra-op/. U.S. EPA (U.S. Environmental Protection Agency). (2007a). Subject: Amended. Imidacloprid. Human Health Risk Assessment. Section 3 Requests for Uses on Peanut, Proso Millet, Pearl Millet, Oat, Kava, Globe Artichoke, Caneberries, Wild Raspberry, and Soybeans. Memorandum from W. Cutchin to S. Jackson. http://www.regulations. gov/fdmspublic/ContentViewer?objectId  090000648023d5e3&dis position  attachment&contentType  pdf. U.S. EPA (U.S. Environmental Protection Agency). (2007b). Revised n-Methyl Carbamate Cumulative Risk Assessment (see Residential Chapter). http://www.epa.gov/oppsrrd1/cumulative/carbamate_fs.htm. U.S. EPA (U.S. Environmental Protection Agency). (2008). d-Phenothrin (Sumithrin®). Addendum to Residential Exposure Assessment. Memorandum from B. Daiss to J. Howenstine and A. Sibold: 326931. Xue, J., Zartarian, V., Moya, J., Freeman, N., Beamer, P., Black, K., Tulve, N., and Shalat, S. (2007). A meta-analysis of children’s handto-mouth frequency data for estimating nondietary ingestion exposure. Risk Anal. 27, 411–420. Zendzian, R. P. (1994). “Dermal Absorption of Pesticides. Pesticide Assessment Guidelines, Subdivision F, Hazard Evaluation, Human and Domestic Animals, Series 85-3,” Health Effects Division, U.S. EPA, Washington, D.C., USA.

Chapter 49

Residential Exposure Assessment: An Overview Jeffrey H. Driver, John H. Ross, Muhilan D. Pandian1, Jeffrey B. Evans2 and Gary K. Whitmyre3 1

infoscientific.com & risksciences.net U.S. Environmental Protection Agency, Office of Pesticide Programs 3 risksciences, LLC 2

49.1  Introduction Following the use of products in and around the home, postapplication chemical exposures to consumers may occur in a variety of microenvironments that correspond to the daily activities in which adults and children engage. These activity patterns may place individuals in contact with a variety of chemicals including pesticides (e.g., dislodgeable foliar residue exposures during gardening, lawn chemical exposures after reentry onto treated turf; and chemical emissions from treated surfaces inside the residence). To understand the potential health significance of these exposures it is necessary to characterize their sources and estimate their magnitude. In response to these needs, efforts have been undertaken to develop methodologies for quantifying pesticide and other chemical exposures in soil, air, food, and water (Cal-EPA, 1994; McKone, 1991, 1993; Ott, 1985; Thompson et al., 1984; Vaccaro et al., 1996; Wallace, 1987, 1989, 1990, 1991; Wallace et al., 1982, 1984, 1985, 1986, 1987a, b, c, 1988, 1989, 1991a, b). The U.S. Environmental Protection Agency (U.S. EPA), for example, in response to the the Food Quality Protection Act of 1996 (FQPA; http:// www.epa.gov/docs/oppfeadsl/fqpa), has been revising exposure monitoring guidelines that emphasize nonoccupational, resi­-dential exposure to pesticides; these guidelines are referred to as “Series 875, Occupational and Residential Exposure Test Guidelines Group B: Post-Application Monitoring Test Guidelines” (http://www.epa.gov/docs/ opptsfrs/OPPTS_Harmonized/875_Occupational_and_ Residential_Exposure_Test_Guidelines; Whitford et al., 1999). The series 875 guidelines provide information and proto­-cols relevant for persons required to submit postapplication exposure data under 40 CFR 158.390 (Fig. 49.1). Generally, these data are required under the Federal Insecticide, Fungicide and Rodenticide Act (FIFRA) when Hayes’ Handbook of Pesticide Toxicology Copyright © 2001 Elsevier Inc. All rights reserved

certain toxicity and/or exposure criteria have been met for a given pesticide (Driver and Wilkinson, 1996). Although at a low level relative to occupational exposures, the major source of chemical exposures for the general population, appears to result from the use of products in and around the home (Hill et al., 1995; U.S. EPA, 1999a; Whitmore et al., 1994). For example, the National Academy of Sciences Committee on Urban Pest Management noted that 5000 health-related incidents involving pesticides were reported as occurring in homes in the United States from 1966 to 1979 (NRC, 1980). More recent data regarding pesticide health-related incidents can be obtained from the American Association of Poison Control Centers (http://www.aapcc.org), the U.S. EPA (http:// www.epa.gov/pesticides), or state regulatory agencies, (e.g., California Department of Pesticide Programs (http://www.cdpr.ca.gov). It is perhaps not surprising, therefore, that the potential health risks associated with exposure to chemicals

Dissipation Studies Dislodgeable Foliar Residue (DFR) Dissipation Study Soil Residue Dissipation (SDR) Study Indoor Surface Residue (ISR) Dissipation Study Measurements of Human Exposure Dermal Exposure (passive dosimetry) Inhalation Exposure Biological Monitoring Other Relevant Data Descriptions of Human Activity Data Data Reporting and Calculations Detailed Product Use Information Figure 49.1  U.S. EPA/OPPTS Series 875, Group B: description of required studies (adapted from ILSI, 1998).

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such as pesticides occurring in and around the home (in air and from surfaces) and from other sources (e.g., consumer products, combustion appliances, and outdoors) are being evaluated much more carefully now than in the past (Driver and Wilkinson, 1996). Pesticides, of course, are just one of many types of chemicals to which humans are exposed in the home. During the past decade and a half, a number of studies, most notably the Total Exposure Assessment Methodology (TEAM) studies sponsored by the EPA, have demonstrated that, for a variety of contaminants, indoor air and other residential pathways are often a more significant source of exposure than corresponding outdoor pathways (Curry et al., 1994; Furtaw et al., 1993; Pellizzari et al., 1987, 1993; Thomas et al., 1993; Wallace, 1993). Indeed, some of the studies found that indoor concentrations of some chemicals were higher than those outdoors and raised serious questions about the relative contribution of indoor sources to the total exposure (Dockery and Spengler, 1981; Melia et al., 1978; Ott, 1985; Spengler et al., 1983). The assessment of potential indoor exposures has also been recognized by industry as a key component in the overall risk evaluation for consumer products (Hakkinen et al., 1991; Hakkinen, 1993). Several studies of potential indoor air exposures to chemicals have been reported to confirm the safety of various consumer products (Gibson et al., 1991; Hendricks, 1970; Wooley et al., 1990).

Figure 49.2  Shematic diagram for nondietary exposure (adapted from ILSI, 1998) Source characteristics Product use patterns and plausible scenarios Human activity patterns Population demographics and stratification

Non-dietary exposure model

Human exposure factors Physico-chemical properties

49.2  Overview of general issues

Residential building factors Temporal and spatial domains

The residential environment should be considered in very dynamic terms. Chemicals that are released into or otherwise enter the residential environment tend to partition into various compartments, either through direct dispersion in indoor air or through adsorption onto surfaces that serve as “sinks” from which material can subsequently be released into the air (Ross et al., 1990, 1991). The amount of a given chemical in each compartment is depleted over time by various mechanisms, including air exchange with the outdoors or with other rooms of the house, and by chemical or physical transformation and/or degradation. Simulation of the behavior of pesticides (and other chemicals) in residential environments can be modeled using the principles of fugacity (Matoba, 1996). Fugacity employs the unit of pressure (Pa), which refers to the external force of a chemical escaping from one compartment or medium to another. For example, there is evidence that particulate contaminants, whether generated inside the residence or brought in from outdoors are adsorbed to surfaces and are later resuspended and recycled within the house after a disturbance (e.g., walking on floors and rugs, sweeping and dusting, and vacuuming; Roberts et al., 1992). A simplistic depiction of the relationship between potential

Uncertainty analysis Model validation

Figure 49.3  Residential exposure assessment: key components (adapted from ILSI, 1998).

sources and exposure pathways in the context of a residential exposure assessment is illustrated in Fig. 49.2. Thus, the residence can be considered an exposure unit containing multiple compartments with which the human receptor can contact. Figure 49.3 illustrates the components of a residential exposure model and Fig. 49.4 illustrates the decision logic associated with construction of a residential exposure assessment. Although inhalation exposure and indoor air quality have received the most attention to date, there are a number of non-inhalation exposure pathways that are likely to be of equal or greater importance for human residential exposures to pesticides and other chemicals. These include potential dermal exposure to dislodgeable chemical residues from surfaces such as floors and carpets or from hard surfaces resulting from the use of formulations for cleaning and disinfection and potential ingestion of surface contaminants

Chapter | 49  Residential Exposure Assessment: An Overview

Hazard Identification and Toxicity Endpoint Selection (e.g., toxicological benchmarks; time-to-effect; exposure/dose metric selection; relevant subpopulation, etc.)

Evaluation of Consumer and Commercial Product Use Information (e.g., application methods, frequencies, locations, rates, formulation types, active ingredient concentrations, post-application activities, consumer/professional survey data, market share data, etc.)

Evaluation of Exposure Monitoring Data (for actual pesticide and/or relevant surrogate chemicals; data limitations; uncertainty and variability; data quality objectives)

Determination of Relevant (Plausible) Aggregate Exposure Scenarios and Related Routes and Pathways of Exposure

Development and Implementation of Screening-Level Deterministic and/or Stochastic Exposure Models

Model Validation and Refinement Figure 49.4  Decision logic and model development/validation process (adapted from ILSI, 1998).

resulting from hand/object-to-mouth activity, particularly in infants and toddlers. Several studies and/or reviews provide examples of noninhalation residential exposures and the complexities involved (Calvin, 1992; CTFA, 1983; Driver et al., 1989; Eberhart, 1994; ECETOC, 1994; Harris and Solomon, 1992; Harris et al., 1992; Turnbull and Rodricks, 1989; Vermiere et al., 1993).

49.3  Lessons learned from key studies Pesticides applied in and around homes by both professional applicators and consumers are used in different ways for different purposes: (1) indoor uses (e.g., floor sprays or foggers for fleas) and outdoor uses (e.g., treatment of wasp nests and ant mounds; use of antimicrobial products in swimming pools); (2) turf uses (e.g., granular applications for control of soil-dwelling insect pests, preemergent and postemergent herbicide sprays) and ornamental uses (e.g., foliar sprays for shrubs); (3) home garden uses (e.g., fungicide dusts for tomatoes); and (4) structural pest control uses (e.g., structural treatment or insecticidal soil barriers to protect against termite invasion). The vast majority of U.S. households use pesticides (Whitmore et al., 1992) and these uses undoubtedly present many opportunities for exposure during intended, label-directed use, misuse, and accidents (Whitmyre et al., 1996). Other sources of indoor exposure to pesticides for the general population may be

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from ambient air, food, water, ambient particles and indoor house dust (Jenkins et al., 1992; Pellizzari et al., 1993; Wallace, 1991, 1993; Whitmore et al., 1994). Residential pesticide monitoring studies have included general surveys of many different pesticides and measurements of air and surface concentrations of specific pesticides after applications of termiticides, crack and crevice or baseboard treatments, total release aerosols or foggers, broadcast applications, and hand-held sprays (Fenske et al., 1990; Racke and Leslie, 1993; Whitmore et al., 1994). These studies typically demonstrate that, after pesticide use, measurable though relatively low-level residues exist in homes and that indoor exposures are often higher than outdoor exposures. Although, in most cases, such exposures are associated with negligible health risks (Whitmore et al., 1994), potential residential exposures to infants and children continue to be the subject of debate and scientific investigation (Berteau et al., 1989; Byrne et al., 1998; Gibson et al., 1998; NRC, 1993; Ross et al., 1990, 1991; Vaccaro et al., 1996; Zweiner and Ginsburg, 1988). However, pesticide biomonitoring of both adults and children demonstrate that absorbed doses from all sources range from fractions of micrograms to single digit micrograms per kilogram of body weight (Adgate et al., 1998; Hill et al., 1995; Krieger et al., 2000, 2001; Vaccaro et al., 1996). The Nonoccupational Pesticide Exposure Study involving about 250 residents of Jacksonville, Florida and Springfield, Massachusetts, clearly demonstrated measurable levels of indoor exposure. Participants in this study carried personal monitors for 24 hours that provided indoor and outdoor measurements of 32 common household pesticides and structural termiticides (Immerman and Schaum, 1990). The indoor air concentrations of these materials exceeded the outdoor air levels by factors even larger than those measured in the earlier TEAM studies on volatile organic compounds (VOCs). It was hypothesized that, with termiticides and other chemicals used outside the home, some material entered the house on soil particles in addition to infiltration in air from treated areas beneath and around the house. It was also noted that the use of “walkoff” rugs in hallways and the practice of removing shoes on entering the house would help to reduce indoor exposure levels (Roberts et al., 1992). In the TEAM studies mentioned earlier, median personal concentrations of VOCs in U.S. residences were found to be 2–5 times outdoor levels and maximum personal concentrations were 5–70 times the highest outdoor levels (Wallace, 1993). The variability in indoor personal exposures probably reflects differences in human activity patterns that bring individuals into contact with chemicals indoors and suggests the importance of specific sources of residential exposures that may not be available to all individuals. Smokers, for example, had benzene exposures 6–10 times higher than those of nonsmokers, individuals wearing freshly dry

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cleaned clothes had significantly higher exposures to tetrachloroethylene (Wallace et al., 1991c), and persons using mothballs and solid deodorizers in the residence had greatly elevated exposures to p-dichlorobenzene relative to those of nonusers (Wallace, 1993). The most recent TEAM study, known as PTEAM, focused on measuring personal exposures to inhalable particles (PM10) of approximately 200 Riverside, California, residents using specially designed indoor sampling devices. A major finding from this study was that personal exposures to particles in the daytime were 50% greater than either indoor or outdoor concentrations (Wallace, 1993). It has been hypothesized that these data suggest that individuals are exposed to a “personal cloud” of particles as they go about their daily activities (Wallace, 1993). Resuspension of household dust through walking in the residence or by vacuuming, cooking, or sharing a home with a smoker lead to significant particle exposures. The recent Valdez Air Health Study conducted in Valdez, Alaska (Wallace, 1993) generally supports the findings of the TEAM studies in terms of the greater importance of individual indoor exposure sources than outdoor sources. In the Valdez study, mean personal concentrations of benzene were roughly 3–4 times higher than outdoor levels, despite the presence of a significant outdoor source of benzene in the community (i.e., a petroleum storage and loading terminal).

49.4  Guidance for residential postapplication exposure assessment methods and data sources for exposure factors The most recent effort to develop guidance for residential exposure assessment methods was initiated by the EPA’s Office of Pesticide Programs in the Standard Operating Procedures for Residential Exposure Assessment (U.S. EPA, 1999a). The passage of the FQPA mandated the EPA to immediately begin routinely addressing nondietary and non-occupational pesticide exposures for the general population. These are exposures that can occur in a residential setting (or other areas frequented by the general population) and that do not occur as part of the diet or as a result of participation in occupational practices. These exposures may include breathing vapors while inside a treated home, exposures to children playing on a treated lawn, or exposures attributable to the mouthing behaviors of infants and children. Before passage of the FQPA, the Agency addressed these kinds of exposures on a case-by-case basis, typically in the “special review” process. The intent of the EPA SOPs is to provide a means for consistently calculating single pathway, screening level exposures and not to provide guidance on other related topics such as aggregate (multisource to a given pesticide) or cumulative (multisource to two or more pesticides with a presumed common mode

of action) exposure assessment. These SOPs are the backbone of the Agency’s current approach for completing initial tier (screening-level) residential exposure assessments. However, the state-of-the-science continues to evolve since the release of the original document in 1997 and the emphasis of industry, as well as of academia and others, has clearly focused on the scientific and policy issues raised by the implementation of the FQPA and the use of the firstgeneration SOPs. Thus, revisions to the SOPs are ongoing to reflect the development of scientific information and the development of refined methods for estimation of potential residential exposures to adults and children. Additional guidance for dermal exposure assessment methods and dermal permeability coefficients for some organic chemicals are contained in the EPA’s dermal exposure assessment guidance document (U.S. EPA, 1992). Given that skin surface area and body weight are closely correlated, total skin surface area to body weight ratios for use in residential exposure assessments have been recommended (Phillips et al., 1993). Another excellent source for methodology and data relevant to consumer product exposure assessments is ECETOC (1994). A number of relevant data sources exist for key variables or factors used in performing residential exposure assessment. Data useful in estimating human exposures (e.g., distributions of body weights and skin surface areas, inhalation rates, and residential occupancy periods) can be obtained from the American Industrial Health Council’s Exposure Factors Sourcebook1 (AIHC, 1995) and the EPA’s Exposure Factors Handbook (U.S. EPA, 1999b), which has recently been updated. Residential “environmental factors” such as air exchange rates have been summarized by Pandian et al. (1993). Human time-activity data in the United States were summarized by the EPA (U.S. EPA, 1991) and compiled in the THERdbASE software (Pandian and Furtaw, 1995), which is available on the Internet at http://www.therd. com. Multiple data sources for time-activity data have been included in EPA’s Consolidated Human Activity Database, which is planned for future release via the Internet.

49.5  Research needs Given that the potential for postapplication exposures largely exists because of product use in and around the home, the need to develop and validate models for prediction of multipathway, multiroute exposures and absorbed dose is evident. Historically, efforts have focused on indoor air and associated inhalation exposures. Jayjock and Hawkins (1993), for example, have explored the complementary roles of indoor air modeling and data development in improving 1

 For the latest version of the Exposure Factors Sourcebook, contact the Update Coordinator, American Industrial Health Council, Suite 760, 2001 Pennsylvania Avenue, N.W., Washington, DC 20006-1807; phone: 202-833-2131.

Chapter | 49  Residential Exposure Assessment: An Overview

the level of confidence in estimates of indoor inhalation exposures. More recently, dermal and incidental ingestion exposures have been the focus of monitoring and modeling efforts (e.g., the Outdoor Residential Exposure Task Force, the Non-Dietary Exposure Task Force, the OP Case Study Group, and Residential Exposure Joint Venture (Zartarian and Leckie, 1998; Zartarian et al., 2000). Multipathway, multiroute modeling efforts for pesticides include the Residential Exposure Assessment Model (REAM), the Stochastic Human Exposure and Dose Simulation model (SHEDS); the Cumulative and Aggregate Risk Evaluation System (CARES), and LifeLine (U.S. EPA, 1999a). The use of real-world data to validate residential exposure models is critical to developing estimates that are more representative than worst-case estimates typically obtained from unvalidated modeling approaches (Whitmyre et al., 1992a, b). Other research activities related to residential exposure assessment currently being sponsored by the EPA include the National Human Exposure Assessment Survey. In addition, the EPA has recently concluded a cooperative agreement, referred to as the Residential Exposure Assessment Project (REAP) with the Society for Risk Analysis and the International Society of Exposure Analysis, to develop a reference textbook describing relevant methodologies, data sources, and research needs for residential exposure assessment. The REAP will complement other EPA initiatives, such as the development of the series 875 guidelines and will facilitate a sharing of information and other resources between the EPA, other federal and state agencies, industry, academia, and other interested parties. Residential exposures to pesticides and other chemicals are estimated by means of either monitoring and/or predictive modeling but, unfortunately, little or no guidance is available for those attempting such estimates. Key areas requiring attention include: Characterization of temporal product use patterns (particularly the likelihood of co-occurrence of more than one product use event) and associated demographic and postapplication activity information relevant to occupants of homes using products. l Source characterization, including emission rates, surface deposition, tranferability to human clothing and skin, and physicochemical factors driving fate and transport processes. l The complex interaction over time of environmental media residue concentrations with humans resulting from variable time-activity patterns that determine subsequent residential exposures (inhalation, dermal, and incidental ingestion). l Identification of the fundamental principles, concepts, and methods for conducting multipathway/multiroute residential exposure assessments, including unique pathways such as incidental dermal exposure to dislodgeable pesticide residues from treated lawns, incidental l



l



l



l



l



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ingestion of contaminated soil particles during gardening, hand-to-mouth transfer by infants and children, and dermal exposure to dislodgeable pesticide residues from carpets and other treated surfaces, and incidental ingestion of postapplication residues in food. Characterization of key human exposure factors (ranges and distributions of factors such as age-specific inhalation rates, product use patterns, and human time-activity data) and residential building factors (distributional data on housing stock type, number and size of rooms, air exchange rates, source emission rates, and sink effects, i.e., adsorption/desorption from various surfaces in the home) that influence residential exposure and dosimetry. Continued development and validation of methods for measuring and modeling indoor chemical fate processes (e.g., volatilization from surfaces and dislodgeable residue kinetics), chemical concentrations in complex matricies (such as house dust) and human intake (e.g., incidental ingestion, inhalation, and dermal exposure). Development and validation of methods for extrapolating from short-term monitoring data to long-term exposure scenarios and for extrapolation of adult monitoring data to children. Continued development and application of methods for quantifying uncertainty and variability (e.g., Monte Carlo methods) in residential exposure (and risk) estimates. The development and use of effective methods for comparing and communicating residential exposure and risk estimates to risk managers and the general public.

References Adgate, J., Quackenboss, J., Needham, L., Pellizari, P., Lioy, P., Shubat, P., Sexton, K. (1998). Comparison of urban versus rural pesticide exposure in Minnesota children. In “Annual Conference of International Society for Environmental Epidemiology (ISEE) and International Conference for Society of Exposure Analysis (ISEA).” July 1998, Vol. 9, No. 4, Suppl., Abstract 920. American Industrial Health Council (AIHC) (1995). “Exposure Factors Sourcebook,”. AIHC, Washington, DC. Berteau, P. E., Knaak, J. B., Mengle, D. C., and Schreider, J. B. (1989). Insecticide absorption from indoor surfaces. In Biological Monitoring for Pesticide Exposure. ACS Symposium Series (R. G. Wang, C. A. Franklin, R. C. Honeycutt, and J. C. Reinert, eds.), Vol. 382, pp. 315–326. Am. Chem. Soc, Washington, DC. Byrne, S. L., Shurdut, B. A., and Saunders, D. G. (1998). Potential chlorpyrifos exposure to residents following standard crack and crevice treatment. Env. Health Perspect. 106, 725–731. California Environmental Protection Agency (Cal-EPA) (1994). CalTOXA™, a Multimedia Total Exposure Model for Hazardous-Waste Sites. Spreadsheet user’s guide, version 1.5. NTIS Publication No. PB95100467. Office of Scientific Affairs, Dep. of Toxic Substances Control, Cal-EPA, Sacramento. Calvin, G. (1992). Risk management case history—detergents. In “Risk Management of Chemicals” (M. L. Richards ed.). Royal Society of Chemistry, United Kingdom.

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Cosmetic, Toiletry and Fragrance Association, Inc. (CTFA) (1983). “Summary of the Results of Surveys of the Amount and Frequency of Use of Cosmetic Products by Women,” Report prepared by ENVIRON Corp.. CTFA, Washington, DC. Curry, K. K., Brookman, D. J., Whitmyre, G. K., Driver, J. H., Hackman, R. J., Hakkinen, P. J., and Ginevan, M. E. (1994). Personal exposures to toluene during use of nail lacquers in residences: description of the results of a preliminary study. J. Expos. Anal. Environ. Epidemiol. 4, 443–456. Dockery, D. W. and Spengler, J. D. (1981). Indoor-outdoor relationships of respirable sulfates and particles. Atmos. Environ. 15, 335–343. Driver, J. H., Konz, J. J., and Whitmyre, G. K. (1989). Soil adherence to human skin. Bull. Environ. Contam. Toxicol. 43, 814–820. Driver, J. H. and Wilkinson, C. F. (1996). Pesticides and human health: science, regulation and public perception. In “Risk Assessment and Management Handbook for Environmental, Health and Safety Professionals” (R. V. Kalluro, S. M. Bartell, R. M. Pitblado, and R. S. Stricoff, eds.). McGraw-Hill, New York. Eberhart, D. C. (1994). Current activities in assessing human exposures to lawn chemicals. In “Workshop on Residential Exposure Assessment, Annual Meeting of the International Society for Exposure Analysis and the International Society for Environmental Epidemiology,” Research Triangle Park, NC. European Centre for Ecotoxicology and Toxicology of Chemicals (ECETOC). (1994). Tech. Rep. 58, Assessment of Non-Occupational Exposure Chemicals, Brussels. Fenske, R. A., Black, K. G., Elkner, K. P., Lee, C., Methner, M. M., and Soto, R. (1990). Potential exposure and health risks of infants following indoor residential pesticide applications. Am. J. Publ. Health 80, 689–693. Furtaw, E. J., Pandian, M. D., and Behar, J. V. (1993). Human exposure in residences to benzene vapors from attached garages. In “Proceedings of International Conference: Indoor Air ‘93,” Helsinki, Finland. Gibson, J. E., Peterson, R. K. D., and Shurdut, B. A. (1998). Human exposure and risk from indoor use of chlorpyrifos. Environ. Hlth. Perspect. 106, 303–306. Gibson, W. S., Keller, F. R., Foltz, D. J., and Harvey, G. J. (1991). Diethylene glycol monobutyl ether concentrations in room air from application of cleaner formulations to hard surfaces. J. Expos. Anal. Environ. Epidemiol. 1, 369–383. Hakkinen, P. J. (1993). Cleaning and laundry products: human exposure assessments. In Handbook of Hazardous Materials, pp. 145–151. Hakkinen, P. J., Kelling, C. K., and Callender, J. C. (1991). Exposure assessment of consumer products: human body weights and total body surface areas to use, and sources of data for specific products. Vet. Hum. Toxicol. 33, 61–65. Harris, S. A. and Solomon, K. R. (1992). Human exposure to 2,4-D following controlled activities on recently-sprayed turf. J. Environ. Sci. Health B27(l), 9–22. Harris, S. A., Solomon, K. R., and Stephenson, G. R. (1992). Exposure of homeowners and bystanders to 2,4-dichlorophenoxyacetic acid (2,4-D). J. Environ. Sci. Health B27(l), 23–38. Hendricks, M. G. (1970). Measurement of enzyme laundry product dust levels and characteristics in consumer use. J. Am. Oil. Chem. Soc. 47, 207–211. Hill, R. H. Jr., Head, S. L., Baker, S., Gregg, M., Shealy, D. B., Bailey, S. L., Williams, C. C., Sampson, E. J., and Needham, L. L. (1995). Pesticide residues in urine of adults living in the United States: reference range concentrations. Environ. Res. 71, 99–108. Immerman, W. W. and Schaum, J. L. (1990). “Nonoccupational Pesticide Exposure Study (NOPES),” NTIS Publication No. PB90-152224, 256 p.

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Report prepared by Research Triangle Institute, Research Triangle Park, NC. Atmospheric Research and Exposure Assessment Laboratory, U.S. Environmental Protection Agency, Research Triangle Park, NC. International Life Sciences Institute (ILSI) (1998). “Aggregate Exposure Assessment” (S.S. Olin, ed.), ILSI Risk Sciences Institute Workshop Report, Washington, DC. Jayjock, M. A. and Hawkins, N. C. (1993). A proposal for improving the role of exposure modeling in risk assessment. Am. Ind. Hyg. Assoc. J. 54, 733–741. Jenkins, P. L., Phillips, T. H., Mulberg, E. J., and Hui, S. P. (1992). Activity patterns of Californians: use of and proximity to indoor pollutant sources. Atmos. Environ. 26, 2141–2148. Krieger, R. I., Bernard, C. E., Dinoff, T. M., Fell, L., Osimitz, T. G., Ross, J. H., and Thongsinthusak, T. (2000). Biomonitoring and whole body cotton dosimetry to estimate potential human dermal exposure to semivolatile chemicals. J. Expos. Anal. Environ. Epidemiol. 10, 50–57. Krieger, R. I., Bernard, C. E., Dinoff, T. M., Ross, J. H., and Williams, R. L. (2001). Biomonitoring of persons exposed to insecticides used in residences. Ann. Occup. Hyg. 45(1), 5143–5153. Matoba, Y. (1996). Simulation of indoor behavior of insecticides applied by various methods. In SP World. No. 24. Sumitomo Chemical Co., Osaka, Japan. McKone, T. E. (1991). Human exposure to chemicals from multiple media and through multiple pathways: research overview and comments. Risk Anal. 11, 5–10. McKone, T. E. (1993). Understanding and modeling multipathway exposures in the home. In “Reference House Workshop II: Residential Exposure Assessment for the ‘90s.” Society for Risk Analysis, 1993 Annual Conference, Savannah, GA. Melia, R. J. W., Florey, C.duV., Darby, S. C., Palmes, E. D., and Goldstein, B. D. (1978). Differences in NO2 levels in kitchens with gas or electric cookers. Atmos. Environ. 12, 1379–1381. National Research Council (NRC) (1980). “Committee on Urban Pest Management.” Nat. Acad. Press, Washington, DC. National Research Council (NRC) (1993). “Pesticides in the Diets of Infants and Children Committee on Pesticides in the Diets of Infants and Children,” Board on Agriculture and Board on Environmental Studies and Toxicology, Commission on Life Sciences. Nat. Acad. Press, Washington, DC. Ott, W. R. (1985). Total human exposure: an emerging science focuses on humans as receptors of environmental pollution. Environ. Sci. Technol. 19, 880. Pandian, M. D., Ott, W. R., and Behar, J. V. (1993). Residential air exchange rates for use in indoor air and exposure modeling studies. J. Expos. Anal. Environ. Epidemiol. 3, 407–416. Pandian, M. D. and Furtaw, E. J. (1995). “THERdbASE: Total Human Exposure Relational Database and Advanced Simulation Environment,” Developed under contract to the U.S. EPA, Office of Research and Development, Environmental Monitoring Systems Laboratory, Las Vegas. Harry Reid Center for Environmental Studies, University of Nevada at Las Vegas, Las Vegas. Pellizzari, E. D., Hartwell, T. D., Perritt, R. L., Sparacino, C. M., Sheldon, L. S., Whitmore, R. W., and Wallace, L. A. (1987). Comparison of indoor and outdoor residential levels of volatile organic chemicals in five U.S. geographic areas. Environ. Int. 12, 619–623. Pellizzari, E. D., Thomas, K. W., Clayton, C. A., Whitmore, R. W., Shores, R. C., Zelon, H. S., and Peritt, R. L. (1993). “Particle Total Exposure Assessment Methodology (PTEAM): Riverside, California Pilot Study,” Vol. I, EPA/600/SR-93/050. U.S. EPA, Research Triangle Park, NC.

Chapter | 49  Residential Exposure Assessment: An Overview

Phillips, L. J., Fares, R. J., and Schweer, L. G. (1993). Distributions of total skin surface area to body weight ratios for use in dermal exposure assessments. J. Expos. Anal. Environ. Epidemiol. 3, 331–338. Racke, K. D., and Leslie, A. R. (eds.). (1993). “ACS Symposium Series 522. Pesticides in Urban Environments: Fate and Significance.” 203rd National Meeting of the American Chemical Society, San Francisco, California, April 5-10, 1992. Published by the American Chemical Society, Washington, DC, ISBN 0-8412-2627-X, 378 p. Roberts, J. W., Budd, W. T., Ruby, M. G., Camann, D. E., Fortmann, R. C., Lewis, R. G., Wallace, L. A., and Spittler, T. M. (1992). Human exposure to pollutants in the floor dust of homes and offices. J. Expos. Anal. Environ. Epidemiol. 1(1), 127–146. Ross, J. H., Fong, H. R., Thongsinthusak, T., Margetich, S., and Krieger, R. (1991). Measuring potential dermal transfer of surface pesticide residue generated from indoor fogger use: using the CDEA roller method. Interim report II. Chemosphere 22, 975–984. Ross, J., Thongsinthusak, T., Fong, H. R., Margetich, S., and Krieger, R. (1990). Measuring potential dermal transfer of surface pesticide residue generated from indoor fogger use: an interim report. Chemosphere 20, 349–360. Spengler, J. D., Duffy, C. P., Letz, R., Tibbets, T. W., and Ferris, B. G. Jr. (1983). Nitrogen dioxide inside and outside 137 homes and implications for ambient air quality standards and health effects research. Environ. Sci. Technol. 17(3), 164–168. Thomas, K. W., Pellizzari, E. D., Clayton, C. A., Perritt, R. L., Dietz, R. N., Goodrich, R. W., Nelson, W. C., and Wallace, L. A. (1993). Temporal variability of benzene exposures for residents in several New Jersey homes with attached garages or tobacco smoke. J. Expos. Anal. Environ. Epidemiol. 3, 49–73. Thompson, D. G., Stephenson, G. R., and Sears, M. K. (1984). Persistence, distribution, and dislodgeable residues of 2,4-D following its application to turfgrass. Pestic. Sci. 15, 353–360. Turnbull, D. and Rodricks, J. V. (1989). A comprehensive risk assessment of DEHP as a component of baby pacifiers, teethers and toys. In “The Risk Assessment of Environmental and Human Health Hazards: A Textbook of Case Studies” (D. J. Paustenbach ed.). Wiley, New York. U.S. Environmental Protection Agency (U.S. EPA) (1991). “Time Spent in Activities, Locations, and Microenvironments: A CaliforniaNational Comparison,” USEPA Publ. No. 600/4-91/006. Office of Research and Development, Environmental Monitoring Systems Laboratory, Las Vegas. U.S. Environmental Protection Agency (U.S. EPA) (1992). “Dermal Exposure Assessment: Principles and Applications,” Publ. No. 600/8-91-011. Exposure Assessment Group, Office of Health and Environmental Assessment, Office of Research and Development, Washington, DC. USEPA, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA). (1999a). “Overview of Issues Related to the Standard Operating Procedures for Residential Exposure Assessment.” Presented to the EPA Science Advisory Panel U.S. EPA, OPP, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA) (1999b). “Exposure Factors Handbook,” USEPA Publ. No. EPA/600/C-99/001. National Center for Environmental Assessment, Cincinnatti. Vaccaro, J. R., Nolan, R. J., Murphy, P. G., and Berbrich, D. B. (1996). “The Use of Unique Study Design to Estimate Exposure of Adults and Children to Surface and Airborne Chemicals,” STP 1287, pp. 166–183. Am. Soc. for Testing and Materials, West Conshohocken, PA. Vermiere, T. G., van der Poel, P., van de Laar, R. T. H., and Roelfzema, H. (1993). Estimation of consumer exposures to chemicals: applications of simple models. Sci. Total Environ. 136, 155–176.

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Wallace, L. A. (1987). “The TEAM Study: Summary and Analysis,” Vol. I,” EPA 600/6-87/002a. U.S. Environmental Protection Agency, Office of Research and Development, Nat. Tech. Information Service, Springfield, VA. Wallace, L. A. (1989). The exposure of the general population to benzene. Cell Biol. Toxicol. 5, 297–314. Wallace, L. A. (1990). Major sources of exposure to benzene and other volatile organic compounds. Risk Anal. 10, 59–64. Wallace, L. A. (1991). Comparison of risks from outdoor and indoor exposure to toxic chemicals. Environ. Health Perspect. 95, 7–13. Wallace, L. (1993). A decade of studies of human exposure: what have we learned? Risk Anal. 13, 135–139. Wallace, L. A., Zweidinger, R., Erickson, M., Cooper, S., Whitaker, D., and Pellizzari, E. (1982). Monitoring individual exposure: measurement of volatile organic compounds in breathing-zone air, drinking water, and exhaled breath. Environ Internal. 8, 269–282. Wallace, L. A., Pellizzari, E., Hartwell, T., Rosenzweig, R., Erickson, M., Sparacino, C., and Zelon, H. (1984). Personal exposure to volatile organic compounds: I. Direct measurement in breathing-zone air, drinking water, food, and exhaled breath. Environ. Res. 35, 293–319. Wallace, L. A., Pellizzari, E., Hartwell, T., Sparacino, C., Sheldon, L., and Zelon, H. (1985). Personal exposures, indoor-outdoor relationships and breath levels of toxic air agents measured for 355 persons in New Jersey. Atmos. Environ. 19, 1651–1661. Wallace, L. A., Pellizzari, E., Hartwell, T., Whitmore, R., Sparacino, C., and Zelon, H. (1986). Total exposure assessment methodology (TEAM) study: personal exposures, indoor-outdoor relationships, and breath levels of volatile organic compounds in New Jersey. Environ. Int. 12, 369–387. Wallace, L. A., Pellizari, E. D., Hartwell, T. D., Sparacino, C., Whitmore, R., Sheldon, L., Zelon, H., and Perrit, R. (1987a). The TEAM study: personal exposures to toxic substances in air, drinking water, and breath of 400 residents of New Jersey, North Carolina, and North Dakota. Environ. Res. 43, 290–307. Wallace, L. A., Pellizari, E., Hartwell, T., Perritt, K., and Ziegenfus, R. (1987b). Exposures to benzene and other volatile organic compounds from active and passive smoking. Arch. Environ. Health 42, 272–279. Wallace, L. A., Pellizari, E., Leaderer, B., Hartwell, T., Perritt, R., Zelon, H., and Sheldon, L. (1987c). Emissions of volatile organic compounds from building materials and consumer products. Atmos. Environ. 21, 385–393. Wallace, L. A., Pellizzari, E. D., Hartwell, T. D., Whitmore, R., Perritt, R., and Sheldon, L. (1988). The California TEAM study: breath concentrations and personal exposures to 26 volatile compounds in air and drinking water of 188 residents of Los Angeles, Antioch, and Pittsburgh, CA. Atmos. Environ. 22, 2141–2163. Wallace, L. A., Pellizzari, E. D., Hartwell, T. D., Davis, V., Michael, L. C., and Whitmore, R. W. (1989). The influence of personal activities on exposure to volatile organic compounds. Environ. Res. 50, 37–55. Wallace, L. A., Nelson, W. C., Ziegenfus, R., and Pellizzari, E. (1991a). The Los Angeles TEAM study: personal exposures, indoor-outdoor air concentrations, and breath concentrations of 25 volatile organic compounds. J. Expos. Anal. Environ. Epidemiol. 1(2), 37–72. Wallace, L. A., Pellizzari, E., and Wendel, C. (1991b). Total volatile organic concentrations in 2700 personal, indoor, and outdoor air samples collected in the USEPA TEAM studies. Indoor Air 4, 465–477. Wallace, L. A., Pellizzari, E., Sheldon, L., Hartwell, T., Perritt, R., and Zelon, H. (1991c). Exposures of dry cleaning workers to tetrachloroethylene and other volatile organic compounds: Measurements

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in air, water, breath, blood, and urine. In “Annual Meeting of the International Society for Exposure Analysis and Environmental Epidemiology,” Atlanta. Whitford, F., Kronenberg, J., Lunchick, C., Driver, J., Tomerlin, R., Wolt, J., Spencer, H., Winter, C., and Whitmyre, G. (1999). “Pesticides and Human Health Risk Assessment: Policies, Processes and Procedures,” Purdue Pesticide Programs, Publication PPP-48. Purdue Univ. Cooperative Extension Service, West Lafayette, IN. Whitmore, R. W., Kelly, J. E., and Reading, P. L. (1992). “National Home and Garden Pesticide Use Survey: Final Report,” NTIS PB92-174739. U.S. EPA, Office of Pesticide Programs and Toxic Substances, Washington, DC. Whitmore, R. W., Immerman, F. W., Camann, D. E., Bond, A. E., Lewis, R. G., and Schaum, J. L. (1994). Non-occupational exposures to pesticides for residents of two U.S. cities. Arch. Environ. Contam. Toxicol. 26, 47–59. Whitmyre, G. K., Driver, J. H., Ginevan, M. E., Tardiff, R. G., and Baker, S. R. (1992a). Human exposure assessment. I: understanding the uncertainties. Toxicol. Ind. Health 8, 297–320.

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Whitmyre, G. K., Driver, J. H., Ginevan, M. E., Tardiff, R. G., and Baker, S. R. (1992b). Human exposure assessment. II: quantifying and reducing the uncertainties. Toxicol. Ind. Health 8, 321–342. Whitmyre, G. K., Driver, J. H., and Hakkinen, P. J. (1996). Assessment of residential exposures to chemicals. In “Fundamentals of Risk Analysis and Risk Management” (V. Molak ed.), pp. 125–141. CRC Lewis Publishers, Boca Raton, Florida. Wooley, J., Nazaroff, W. W., and Hodgson, A. T. (1990). Release of ethanol to the atmosphere during use of consumer cleaning products. J. Air Waste Management Assoc. 40, 1114–1120. Zartarian, V. G. and Leckie, J. O. (1998). Dermal exposure: the missing link. Environ. Sci. Technol. March 1, 134–137. Zartarian, V. G., Ozkaynak, H., Burke, J. M., Zufall, M. J., Rigas, M. L., and Furtaw, E. J. (2000). Amodeling framework for estimating residential exposure to and dose of chlorpyrifos via dermal residue contact and non-dietary ingestion. Environ. Health Perspect. 108, 505–514. Zweiner, R. J. and Ginsburg, C. M. (1988). Organophosphate and carbamate poisoning in infants and children. Pediatrics 81, 121–126.

Chapter 50

Modeling Dietary Exposure with Special Sections on Modeling Aggregate and Cumulative Exposure Barbara J. Petersen Exponent , Inc. Washington, DC

50.1  Overview Pesticide residues can be found in food as a result of the use of the pesticide to protect the crop from pests. Residues are also found on surfaces in the home and garden following the use of pesticides for a variety of different purposes (pest control, lawn treatments, etc.). Pesticide residues can be found in the air in the home as a result of volatilization from water or volatilization during and after the application of the pesticide. Thus each of these routes must be considered in estimating exposure of consumers to a pesticide or pesticides. For some pesticides it is possible to estimate total exposure (regardless of the route of exposure) by using biomonitoring techniques. These techniques rely on analyses of urine (or other body fluids) to determine the levels of the pesticide and/or its metabolites. The results can then be used to estimate the person’s exposure by combining them with knowledge about the chemical’s absorption into the body along with information about the metabolism of the chemical once it is absorbed. This approach is not feasible in many situations – either because the pesticide has not yet been approved for use so there is no exposure or because there is insufficient information about the relationship between exposure and urinary metabolites. Also, biomonitoring methods do not identify the source of the exposure. Therefore, in many situations other assessment methods are necessary to identify the potential importance of a specific route relative to other pathways of exposure. These methods are generally referred to as indirect methods. Ideally, the selected method should identify the proportion of exposure that is due to each route, for example, that proportion that is due to oral exposure, dermal exposure, or inhalation of the pesticide. In many cases exposures from Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

more than one source need to be considered at the same time. For example, oral exposure can arise as a result of residues in the diet or from other pathways such as toddler hand-to-mouth activity. Typically, indirect exposures are estimated for each route and then combined in order to estimate total exposure. Again, the three potential routes are oral, dermal, and inhalation. Oral exposures occur as a result of pesticide residues in food and/or as a result of residues that are in other media that are transferred to the hand and thence to the mouth. Dermal exposures occur as a result of pesticides in air, in water, or on surfaces that are touched and subsequently absorbed through the skin. Inhalation exposures occur as a result of the pesticide’s presence in the air that individuals breathe and absorption into the body through the lung. Indirect estimates of consumer exposure to a pesticide (or groups of pesticides) require data on the levels in the media (food, water, air, etc.) as well as estimates of the amount of food consumed, extent of dermal exposure (skin area), and inhalation rates. Exposure assessors have developed some more or less standard terminology in order to facilitate the conduct and understanding of exposure analyses. Typically, dietary exposure estimates consider only food as the source of exposure, while aggregate exposure estimates include all exposures to the same chemical, for example, for exposures via multiple pathways including inhalation, dermal, and oral routes. Cumulative exposure assessments estimate exposure to multiple chemicals via multiple pathways, for example, all of the potential routes for a single pesticide or all of the routes for a group of pesticides. Dietary, cumulative, and aggregate exposure methodologies are discussed in this chapter along with examples of the algorithms and data that can be employed for each method. For each route, methods have been developed that 1099

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are highly refined and require extensive data sets that provide more reliable estimates as well as less refined methods based on relatively little data. In many situations the desired data are not available and it is necessary to make a rough estimate using the available data and to include reasonable (or “worst-case”) assumptions. Methods that use default assumptions or simplistic methodologies generate worst-case exposure and risk estimates that are extreme and are generally considered to be screening estimates. Such simplified approaches do not permit any discrimination between factors that are likely to reflect risks that really exist and those that are truly hypothetical. If sufficiently conservative assumptions are made the assessment will surely define the potential upper estimates of exposure. This chapter focuses on methodology to estimate dietary exposures and provides options that range from worst-case screening assessments to much more refined and accurate assessments. Separate sections describe techniques that can be used to conduct aggregate and cumulative assessments. The methods for estimating cumulative exposure assume that it has been determined that the chemicals cause toxicity by a common mechanism. While the toxicity may be caused by the same mechanism, it is not necessary to assume the same potency for each compound. Rather, exposures can be adjusted to reflect the relative potency of each chemical. Figure 50.1 provides a schematic overview of an aggregate and cumulative assessment process. There are several features that are required in any exposure assessment. Examples of common features include the ability to provide estimates of exposure that are statistically representative of the population to be evaluated, the ability to consider the impact of various scenarios and data, and the ability to provide documentation that is suitable for regulatory decisions and/or to support peer review.

Wherever possible it is desirable to be able to estimate exposures for a range of subpopulations (such as different age/sex groups) as well as for the overall population. In the case of aggregate or cumulative exposure, it is also necessary to be able to estimate exposure to single or multiple compounds, respectively, for relevant time periods. The goal of aggregate and cumulative exposure assessments is to characterize the exposure of the population of concern (e.g., adults, toddlers) and to identify the variability and uncertainty associated with that exposure. It is also desirable to be able to estimate exposures over different time periods—for example, for a single day, a month, a year, or even a lifetime. The exposures are characterized by estimating the level of chemical uptake via ingestion, inhalation, and/or dermal absorption of the substance over various time periods. In practice, the time periods over which exposures are typically evaluated include daily/acute, short-term (1–7 days), intermediate-term, and chronic (up to 1 year) time periods. The time periods of interest are based on the toxicity profile of the chemical (and according to existing regulatory requirements). Exposure assessments can be useful to identify the potential importance of a specific route relative to other pathways of exposure. That is, the method should identify the proportion of exposure that can result from oral, dermal, or inhalation or a combination of these routes. In many cases, exposures from more than one source need to be considered. For example, oral exposure can arise as a result of residues in the diet or from other pathways such as toddler hand-to-mouth activity. The methods should allow the user to aggregate exposures as appropriate for the scenarios under consideration. Aggregation may be relevant to one chemical contained in one product that has multiple routes of exposure. For example, a compound might be used exclusively in a lawn

Dietary assessment

USDA PDP Multiple compounds Residue distributions

USDA CSFII Demographic, social, geographical characteristics & food consumption patterns

Non-dietary assessment Residential pesticide use patterns Residue data Residue files

Processing factors Other food residue data

Pesticide use patterns

Other residue adjustments

Residue files

Dietary exposure estimates

Cumulative risk estimates

Toxicological profiles

Figure 50.1  Components of an aggregate and cumulative risk assessment.

Non-dietary exposure estimates

Aggregate exposure estimates

Chapter | 50  Modeling Dietary Exposure with Special Sections on Modeling Aggregate and Cumulative Exposure

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80 Dietary

70

Use B Use C

Exposure (ug/kg)

60

TOTAL

50 40 30 20 10 0 1

31

61

91

121

151

181

211

241

271

301

331

361

Day Figure 50.2  Hypothetical multiuse exposures.

product that is applied by the homeowner. In this case, the person applying the compound may be exposed by dermal and inhalation routes while applying the compound and using the lawn after the treatment. Aggregation might also be relevant to one chemical contained in multiple products and with multiple exposure routes. For example, if a chemical is used on crops and also as a termiticide, possible routes of exposure are oral (from residues on food), dermal (from contact with treated surfaces in the home), and inhalation (from residues in the air around the home). For cumulative exposure assessments, the methods need to simulate situations that encompass exposure to more than one chemical, with multiple uses or sources and multiple exposure routes (Figure 50.2). Cumulative assessments should be limited to chemicals that have a similar mode of action toxicologically. This requires that the relative toxicity of the various chemicals included in the analysis be quantitatively specified. It is also necessary that the exposure methods account for temporal and spatial considerations. The temporal aspect should consider when exposures occur simultaneously. For example, the method should account for overlap of exposures based on product usage information and chemical degradation. The spatial or regional aspect should be included in the exposure methodology. For example, the types of contaminants encountered in a home in Florida may be very different from those found in a home in northern Maine. In summary, exposure estimates must be able to assess exposures that are specific to both time and location.

50.2  Exposure models 50.2.1  General Exposure Model Exposure assessment models can be designed for a specific route or for a specific source of exposure. Routes of

exposure to pesticides can include oral, inhalation, and dermal pathways. Exposures can result from the treatment of foods, homes, schools, workplaces, yards, gardens, parks, golf courses, etc. All exposure models incorporate the following variables in one form or another:

contact  residue  exposure

The “contact” function refers to the amounts of foods eaten for dietary exposure and/or breathing rates/activity patterns, etc. for nondietary exposures. The “residue” function refers to the concentration of the pesticide or pesticides on the foods that were consumed for dietary exposure and to levels in other media (air, water, surfaces) for other routes of exposure. Each function may have multiple parameters and can be represented by worst-case estimates or by refined statistically representative data sets. To assess the dietary aggregate or cumulative exposure, three types of data for each product or use are required: 1. Information about how the chemical enters the media. For example, for pesticides used in the home, information that is needed includes use patterns of products of interest, frequency of application, and amount of product applied. 2. Environmental concentration data on relevant days. Again, for a pesticide used in the home, environmental data are needed before, during, and after treatment (residue factors). 3. Exposure factors such as food consumption information, body weight, breathing rate, and activity patterns (contact factors). Selection of the most appropriate methodology for estimating exposure depends on the intended application or purpose for the exposure assessment, the available data and related resources, and the relative importance of each route of exposure. The exposure route or pathway plays a critical

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role in choosing the most appropriate exposure algorithms for use in the assessment. The appropriate algorithms need to reflect the manner in which the chemical is contacted. The contribution of the diet to exposure must include estimates of consumption of each food as well as the residues in each food (including estimates of the proportion of food that contains residues). The impact of cooking and processing and the likelihood that the water supply also contains residues should be considered. Likewise, factors must be selected for aggregate exposure assessments. For example, if exposure to residues from an indoor treatment is to be determined, the exposure algorithms must account for residues in the air (for the inhalation route) as well as residues deposited on the carpet (these residues will in turn be used to estimate ingestion via hand-to-mouth and dermal routes). Parameters that are not independent need to be treated differently. For example, high consumers of one food may be low consumers of another food, and such relationships must be included in the assessment. In each exposure analysis, routes should be “linked” if exposures are dependent. Linking should not aggregate scenarios; rather, it should assume that exposures occur simultaneously (if they do or could) and should not assume that exposures occur simultaneously (if they do not or are unlikely to occur at the same time). Techniques for considering each of these factors are presented in the following.

50.2.2  General Dietary Model The equation to calculate dietary exposure, is deceptively simple:

exposure  C  R

(1)

where C is the amount of the food consumed and R is the concentration of chemical in the food. The estimate of consumption can include both quantity and frequency of consumption. This equation can be modified to incorporate other parameters as necessary. In a detailed exposure assessment, the basic algorithm [Eq. (1)] is modified to allow the analyst to better match the available data to the toxicity profile of the chemical. Specifically, the analyst will utilize much more extensive data that permit refining who (e.g., toddlers, all adults, adult females) is exposed and under what conditions (e.g., exposure time). For example, exposure can be estimated for an average or median individual or the exposure can describe the distribution of potential exposures for all the individuals in a specific population. Additionally, the analyst must decide the appropriate period of exposure time that is relevant to the toxicity profile to include in the exposure estimate.

50.2.2.1  Populations The population for which exposure is to be estimated depends on the toxicity and use patterns of the pesticide

as well as regulatory requirements. Typically, exposure estimates will be needed for a representative sample of the U.S. population as well as for a range of subpopulations (e.g., toddlers, the elderly, adolescents, females of childbearing age). The U.S. population can be represented by a subsample drawn from the U.S. Census. The U.S. Department of Agriculture (USDA) and the U.S. National Center for Health Statistics (NCHS, 1992–2006) have developed sample frames for the national food consumption surveys that are representative of the U.S. population. The surveys are sufficiently large and contain sufficient demographic information to allow more than 25 subgroups of the population to be evaluated. Demographic data are also available for the USDA and NCHS reference populations that can be used for nondietary assessments as well. The NCHS has created sample frames as a part of each National Health and Nutrition Examination Survey (NHANES). NHANES includes national food consumption surveys that are designed to provide detailed information about foods that are consumed by the U.S. population. Data are released on a periodic basis (currently NHANES data are publically available for years 2005–2006). NCHS has developed a statistical weight for each individual that can be applied to develop estimates of exposure for the U.S. population. The demographic variables for each individual can be used in part to support the selection of parameters for aggregate and cumulative exposure assessments as well as to estimate exposure due to residues in the food supply. The U.S. Environmental Protection Agency (EPA) has developed science policies (and associated regulations) for conducting exposure assessment for pesticides under the U.S. Food Quality Protection Act. Harmonization of these policies between U.S. and international agencies has taken place through the Codex Alimentarius Committee on Pesticide Residues (CCPR) as well with Canada and other countries. Policies have been developed that describe the data required and the degree of conservatism that are appropriate for different scenarios. These policies are designed to provide regulators with the information that is required to make decisions that fully protect public health and that identify sensitive subpopulations or population with unusual exposures.

50.2.2.2  Exposure Period Generally, most models use the calendar day as the default unit of time for calculating human exposure to one or more chemicals. All reporting periods longer than a day can then be “created” by combining sequential daily exposures to an individual that are summed and averaged over the number of days included in the reporting period to provide an average daily exposure for that individual over the time duration specified in the analysis. The World Health Organization (WHO) through its expert committees Joint FAO/WHO Expert Committee on Food Additives and Joint

Chapter | 50  Modeling Dietary Exposure with Special Sections on Modeling Aggregate and Cumulative Exposure

Food/WHO Meeting on Pesticide Residues often estimates weekly exposures for contaminants. Those weekly estimates are then compared to the weekly estimate of acceptable intake levels.

CalendexTM analysis

User defines analysis parameters Launch analysis

50.2.2.3  Screening Models Exposures can result from the treatment of foods, homes, schools, workplaces, yards, gardens, parks, golf courses, etc. Although it is theoretically possible that all of these locations could be treated with the same pesticide, it is extremely unlikely that such a scenario would in fact occur. Nonetheless, an estimate of exposure can be computed by assuming that simultaneous exposures from all of the treatments actually occur. Estimates that make such extreme assumptions are usually called screening or range-finding estimates and have the advantage of being simple and inexpensive (because they do not require a lot of data) to conduct. More realistic estimates, incorporating the probability of exposure, are derived using more complicated, resource-intensive approaches. Screening models also usually estimate a single value (e.g., a mean, median, or maximum). This single exposure metric (often called a point estimate of exposure) for a pesticide is determined by using one number to represent the residue (i.e., concentration of the chemical) and one number to represent contact with the chemical by that population. The most basic models combine data on average contact and average concentration levels of the substance to estimate average exposure. It is also possible to estimate maximum exposures by assuming all food contains the maximum permissible levels and that extremely high amounts are consumed. In estimating average exposure using point estimates, the arithmetic mean is most commonly used; however, if the distribution of the parameter of interest is known to be skewed (as is typical of food consumption data), use of the geometric mean or median (or 50th percentile) concentration is more appropriate (Mosteller and Tukey, 1977).

50.2.3  Probabilistic or Monte Carlo Models Probabilistic or Monte Carlo assessments utilize both contact level distributions (amount of food consumed, time of residence in the home, etc.) and residue distributions. Contact levels vary among individuals. Children typically consume more food per unit body weight than adults. Also, the total dermal contact for children may be larger than the dermal contact level for adults because children may have more active contact with the treated surface. Similarly, residue levels present in the residential environment also vary. Variations in the contact and chemical concentrations produce potential variations in the resulting exposure distributions (Figure 50.3).

1103

i = Nind?

No

Individual (i = 1 to Nind)

No i=i+1

CSFII demograph

Match demographic parameters?

Yes Iteration ( j =1 to Nn)

Yes Compute exposure

Yes

j = Nn? No j=j+1

Compute means, distributions, etc. by exposure route + total exposure

Figure 50.3  General sequence of probabilistic exposure analyses.

50.2.4  Combination of Screening and Probabilistic Models Most analysts use a combination of worst-case estimates for some parameters and more realistic estimates for other parameters. Typically, more resources are devoted to refining the parameters that have the biggest impact on the exposure assessment. WHO is updating its principles and guidances for conducting risk assessments. Recently, the WHO Exposure Assessment Guidelines were published (2008) and are available on the WHO website (http://whqlibdoc.who.int/ publications/2008/9789241597470_eng.pdf).

50.2.5  Contact Factors for Dietary Exposure Model: Food Consumption There have been numerous food consumption surveys conducted to estimate the intake of nutrients. Such surveys have been conducted at periodic intervals in many countries. The methods used to conduct dietary surveys are in a continuing state of development and refinement. The survey instruments and procedures to collect and analyze

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samples are continually being improved, and new compounds are being added to reflect current priorities. The optimal approach for assessing dietary exposures has been debated for years in the United States as well as in other countries (FASEB, 1988). The most appropriate survey for use in an exposure assessment must be determined. In selecting a survey it is important to evaluate the following: When were the data collected and are current dietary practices similar enough to be relevant? l Were appropriate populations and subgroups of the population surveyed? l Were data collected during all seasons? l Were the foods of interest included in the survey? l Was the quantity of each food estimated? l

The available information about the survey methodology and the resulting data available from a particular survey are important determinants of the usefulness of the survey for estimating exposure using the data.

50.2.6  Common Types of Food Consumption Surveys There are four broad categories of food consumption data: (1) food supply surveys (market disappearance), (2) household or community inventories, (3) household food use data, and (4) surveys of individual food consumption patterns.

50.2.6.1  Food Supply Surveys Food supply surveys, which are called food balance sheets (FBSs) by the Food and Agriculture Board of the United Nations, are conducted on a countrywide basis each year by almost every country in the world (FAOSTAT, 2007). These surveys provide data on food availability or disappearance rather than actual food consumption. Typically, the data are collected for the entire country and a per capita estimate is derived by dividing by the number of individuals in the country. The data are typically provided for agricultural commodities. These surveys provide a rough estimate of the amounts of foods consumed by the country’s population. Food supply survey results are useful for setting priorities, analyzing trends, developing policy, and formulating food programs. For some countries, food supply data are the only accessible data that represent the country’s food consumption. Although other types of data are generally better for risk assessment, food supply surveys can be particularly useful when there is a need to compare potential risks in different populations. The FBSs are useful for this because similar methods are used around the world. Food supply surveys also have been useful in some epidemiological studies (Sasaki and Kestelhoot,

Hayes’ Handbook of Pesticide Toxicology

1992). FBSs describe a country’s food supply during a specified time period. Mean per capita availability of a food or commodity is calculated by dividing total availability of the food by the total population of the country. FBSs, published by the FAO, describe the food supply in countries on all continents. European FBSs are also prepared by the Organization for Economic Cooperation and Development and the Statistical Office of the European Communities. Food supply data in the United States are developed by the USDA’s Economic Research Service. There are some limitations in the use of FBSs to estimate exposures. First, waste at the household and individual levels usually is not considered. Therefore, exposure estimates based on food supply data are higher than estimates based on actual food consumption survey data: The magnitude of the error depends on the quantity of waste produced. Perhaps more importantly for exposure assessment, users of foods cannot be distinguished from nonusers. Therefore, individual variations in exposure cannot be assessed, nor can exposure of potentially sensitive subpopulations be estimated. Finally, food availability is usually reported in terms of raw agricultural commodities. Processed forms of foods are usually not considered, and there is no way to distinguish the use of foods as ingredients. Nonetheless, these data allow assessments to be conducted at the international level using comparable information.

50.2.6.2  Household Inventories Household surveys generally can be categorized as (1) household or community inventories or (2) household or individual food use. Inventories are accounts of what foods are available in the household. What foods enter the household? Were they purchased, grown, or obtained some other way? What foods are used up by the household? Were they used by household members, guests, and/or tenants? Were they fed to animals? Inventories vary in the precision with which data are collected. Questionnaires may or may not ask about forms of the food (i.e., canned, frozen, or fresh), source (i.e., grown, purchased, or provided through a food program), cost, or preparation. Quantities of foods may be inventoried as purchased, as grown, with inedible parts included or removed, as cooked, or as raw. Such data are available from many countries including Germany, the United Kingdom, Hungary, Poland, Greece, Belgium, Ireland, Luxembourg, Norway, and Spain (DAFNE II, 1995; Trichopoulou and Pagona, 1997), Australia/NZ (2009), Japan (2005), and China (2006).

50.2.6.3  Household Food Use Data Food use studies, usually conducted at the household or family level, often are used to provide economic data for policy development and planning for feeding programs.

Chapter | 50  Modeling Dietary Exposure with Special Sections on Modeling Aggregate and Cumulative Exposure

Survey methods used include food accounts, inventories, records, and list recalls (Pao et al., 1989). These methods account for all foods used in the home during the survey period. This includes foods used from what was on hand in the household at the beginning of the survey period and foods brought into the home during the survey period. Although household food use data have been used for a variety of purposes, including exposure assessment, serious limitations associated with data from these surveys should be noted. Food waste often is not accounted for. Food purchased and consumed outside the household may or may not be considered. The household members who did and did not consume a particular food cannot be distinguished, and individual variation cannot be determined. Exposures by subpopulations based on age, gender, health status, and other variables for individuals can only be estimated based on standard proportions or equivalents for age/gender categories.

50.2.6.4  Individual Consumption Studies Individual exposure studies provide data on food consumption by specific individuals. Methods for assessing food exposures of individuals may be retrospective (e.g., 24h or other short-term recalls, food frequencies, and diet histories), prospective (e.g., food diaries, food records, or duplicate portions), or a combination thereof. The most commonly used studies are those that use the recall or record method and the food frequency method. For example, national dietary surveys were conducted in Australia in 1983 for adults, in 1985 for school children, and in 1995 for the entire population. The U.S. NHANES survey collects retrospective, prospective, and food frequency data from its respondents (NHANES, 2003–2006). Food consumption data are available from food frequency surveys for some populations (ANZFA, 1997–2006). Food recall studies are used to collect information on foods consumed in the past. The unit of observation is the individual or the household. The subject is asked to recall what foods and beverages he or she or the household consumed during a specific period, usually the preceding 24 h. Because this method depends on memory, foods are quantified retrospectively, often with the aid of pictures, household measures, or two- or three-dimensional food models. Recalls have been used successfully with individuals as young as 6 years of age, and interviewer-administered recalls are usually the method employed for populations with limited literacy or for individuals whose native language is not English. When individuals are not available for an interview or are unable to be interviewed due to age, infirmity, or temporary absence from the household, surrogate respondents often are used (Samet, 1989). The main disadvantage of the recall method is the potential for error due to faulty memory of respondents. Items that were consumed may be forgotten or the respondent may recall items

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consumed that actually were not consumed during the time investigated. To aid recall memories, the interviewer may probe for certain foods or beverages that are frequently forgotten, but this probing also has been shown to introduce potential bias by encouraging reporting of items not actually consumed. Food record/diary surveys collect information about current food exposure by having the subject keep a record of foods and beverages as they are consumed during a specific period. Quantities of foods and beverages consumed are entered in the record usually after weighing, measuring, or recording package sizes. Bassett et al. (2008) obtained recall information from respondents as they left a restaurant, including collecting the respondent’s receipt for the meal and conducting an interview as to whether everything was consumed or not (Bassett et al., 2008). Occasionally, photographs or other recording devices are employed. The results of surveys conducted using short-term recalls or food diaries are generally the most useful source of data to estimate exposure to pesticides. Data from these surveys can be used to estimate either acute or chronic exposure. Averages and distributions can be calculated, and exposure estimates can be calculated for subpopulations based on age, gender, ethnic background, socioeconomic status, and other demographic variables, provided that such information is collected for each individual. Food frequency questionnaire (FFQ) surveys typically allow qualitative estimates of exposure. A FFQ or checklist is used to determine the frequency of consumption of the foods of interest. Subjects indicate how many times each day, week, or month they usually consume each food. Occasionally a semi-quantitative FFQ survey is utilized that estimates the amounts consumed by having respondents indicate whether their usual portion size is small, medium, or large. Foods consumed away from the home can represent a significant proportion of the diet (Lin, 1999). Therefore, regardless of the method or the geographic locale it is important to collect information for foods consumed away from the homes as well as at home. For the U.S. population, the data collected in one of two large national food consumption surveys conducted by the U.S. government, (1) the Nationwide Food Consumption Survey conducted by the USDA, beginning in 1935, or (2) the National Health and Nutrition Examination Survey undertaken by the U.S. Department of Health and Human Services, NCHS, beginning in 1971, are appropriate for most exposure assessments. The two surveys have now been combined into a single survey that is an ongoing survey. Results are released at periodic intervals and can be combined to enlarge the number of subjects. Both surveys employ multistage area probability sampling procedures to obtain a sample that is representative of the population. The NHANES surveys are repeated at periodic intervals and are publicly available.

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50.2.7  Estimating Residues in Foods Estimates of the residues in foods can be made using a single fixed (deterministic) value, by using empirical distributions of residues, or by using common statistical distributions as a representation of the residues. Often an exposure assessment uses single fixed values for some foods and distributions for others. When a distribution is created from other distributions, Monte Carlo sampling is used to choose a value from each distribution independently from one another, and then these values are combined (e.g., multiplied or added together, depending on the nature of the built-up composite function as specified by the user). In other situations, the residues may need to be handled as dependent variables. Residue data can come from a variety of sources, including controlled research studies and government and industry monitoring programs. The types of residue data used in dietary risk assessments include the following. 1. Published tolerances (listed in the U.S. Code of Federal Regulations (CFRs) by pesticide/crop): The published tolerance levels are “worst case” in that residues are at the highest allowable level. 2. Controlled field trial data: The levels estimated from these data are somewhat less conservative since the data were obtained from controlled studies in which residues were measured following treatment of the field according to worst-case label conditions. The residues are somewhat more realistic than the tolerance levels, but still reflect the maximum use rate and minimum pre-harvest intervals. 3. Monitoring data: Monitoring data, such as the FDA TD and the USDA PDP (which are described later), are the most realistic since they reflect actual consumer practices. 4. Effects of processing and cooking: All three types of data can be improved by including the results of processing and cooking. Studies can be conducted that measure pesticide concentrations in the food before and after processing and/or cooking. A processing factor is derived and then applied to the levels estimated in the raw agricultural commodity. 5. Proportion of crop treated: In the case of published tolerance and controlled field trial data, additional information about the proportion of the crop that could be treated may provide a more realistic exposure assessment. There are three common approaches that are appropriate for collecting monitoring data on concentrations of chemicals in food: (1) duplicate diets, (2) market basket or representative sampling, and (3) controlled experimentation. The USDA monitors residue levels in selected fruits and vegetables and meat and poultry products. The FDA monitors residue levels in all other foods. California,

Hayes’ Handbook of Pesticide Toxicology

Florida, and a number of other states also have monitoring programs. Depending on the specific U.S. monitoring program, foods or commodities may be sampled at the point of entry to the country, at the farm gate, at the food processing plant, or at the retail level. The FDA conducts estimates of exposure to some pesticides through its Total Diet Study (TDS). The TDS, sometimes called the market basket study, is an ongoing FDA program that determines levels of various contaminants and nutrients in foods. From this information, dietary intakes of those analytes by the U.S. population can be estimated using some “model” estimates of consumption patterns. Additional information about the TDS is available on the U.S. FDA website (US FDA, 2006). It is also possible to download the analytical results (data summaries as well as individual results for each food) from mid-1991 to the present. Other data on pesticide residues in the food supply are available through the Pesticide Data Program (PDP), a national pesticide residue database program. Through cooperation with state agriculture departments and other federal agencies, PDP manages the collection, analysis, data entry, and reporting of pesticide residues on agricultural commodities in the U.S. food supply, with an emphasis on those commodities highly consumed by infants and children. Additional information is available at the PDP website: http://www.ams.usda.gov/AMSv1.0/ams.fetchTemplateData.do?templateTemplateC&navIDPestici deDataProgram&rightNav1PesticideDataProgram&top Nav&leftNavScienceandLaboratories&pagePestic ideDataProgram&r. Controlled experiments are sometimes used to determine the likely levels of pesticides in foods produced under specified conditions and of relevant metabolites. Field trials, conducted under the worst-case conditions of the level, are conducted for virtually all pesticides. Data from these trials are considered to be representative of the extreme upper limit of potential residues. Analyses of foods collected in supermarkets are more likely to be representative of the residues encountered by consumers.

50.2.8  Considerations for Cumulative Exposure Assessments Cumulative exposure assessments must be designed to allow the estimation of the simultaneous exposure to compounds that have the same mechanism of toxicity. Therefore, methods used to estimate exposure to multiple chemicals need to adjust the detected residue levels of each of the chemicals considered by “relative toxicity factors” that reflect the toxicity levels of these chemicals relative to a “standard” chemical. A total adjusted residue is derived for each sample by summing the adjusted residue values for that commodity. An exposure assessment is then conducted using these total adjusted residues. This approach is based

Chapter | 50  Modeling Dietary Exposure with Special Sections on Modeling Aggregate and Cumulative Exposure

on concepts proposed by the National Academy of Sciences for the assessment of joint exposure to organophosphate pesticides and is similar to that followed by the EPA in the case of dioxin-like compounds. This adjustment can be done outside of the exposure assessment model, and a single residue can be entered into the analysis. However, generally some technique is needed to allow a determination of the relative contribution of each chemical to the exposure assessment. Additional information is available from the EPA website: http://www.epa.gov/pesticides/cumulative/.

50.2.9  Consideration for Aggregate Exposure Assessments Typically, more realistic estimates are required for aggregate exposure assessments. Worst-case approaches compile too many conservative values to provide any meaningful evaluation. Types of information that need to be considered include applicator and postapplication exposures that incorporate information about the likelihood that each treatment and contact occur. Probabilistic methods are used to determine whether a given pesticide is applied in a specific household, the application dates of pesticide treatments in the home, the amount of residue uptake per unit of contact, the level of contact by each individual that results in the uptake of physical chemical residues, and other relevant parameters. Chemical exposure by each individual in the sample population realistically can be estimated repeatedly using Monte Carlo analysis that specifies a new set of daily food consumption data, treatment schedules, contact schedules, and residue concentrations with each iteration. Aggregate exposure estimates are improved by including information about the following: Percentage of households treated (e.g., 75% of the households in the United States use flea and tick products) l Regions of treatment (e.g., treatment in Florida occurs all year round, foods from Latin America are untreated) l Type of applicator (e.g., the product is sold only to homeowners) l Frequency of application: typical intervals between when the pesticide is applied and the consumer is exposed l Exposure/biomonitoring studies: As noted previously, sometimes the exposure to a specific pesticide can be estimated directly from biological samples. Such data may be point estimates or empirical or parametric distributions of exposure.

l

50.2.10  Estimating Aggregate and Cumulative Exposures Using a Calendar Model The EPA has developed standardized approaches for estimating aggregate and cumulative exposures. Typical analyses

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are available on the EPA website: http://www.epa.gov/pesticides/cumulative/2006-op/index.htm. A calendar model can be used to account for the variation in application and exposure. This is important since not all products are applied on the same day or used with the same frequency. Information about the frequency of use or application is very important. This information may be obtained from market use data. If market use data are not available, the information on the product label can be used to improve the estimates. Typically, product labels specify minimum frequencies of application. It is also possible to apply reasonable seasonal assumptions that are made based on professional experience. It is important to recognize that residues from various treatments might overlap. For example, a professional applicator may treat a home for cockroaches on March 1. On June 1, the homeowner may spray the lawn with the same active ingredient, and foods may be treated at different times during the growing season. The family dog may be treated for fleas on August 1 using the same active ingredient. The question might then arise, “What is the exposure to a child on September 7?” A realistic model will not assume the worst case for each of these scenarios and simply add the residues together. Instead, a good model will account for such overlap based on the probability of occurrence of such overlap and chemical degradation. Figure 50.4 presents a graphical representation of how a calendar model determines the available residue per unit of contact on September 7 (day 250 of the year) from the three treatments previously described. In summary, a calendar model Uses the probability that individual exposures occur around specific dates l Calculates exposure for individual chemical uses and exposure routes l Combines the exposure-probability distributions for individual uses using Monte Carlo sampling techniques. l

The calendar model is able to estimate the available residue value on the day of exposure by first computing the number of elapsed days between the application day and the day of exposure for each use of each chemical. A degradation function can then be used to adjust the levels of chemical residues present. This approach is often not feasible to estimate food residues, and assumptions must be made about which residues are present at a given time. Estimates of residues by season are available for many chemical and crop combinations. For each day of the year, the calendar model can combine exposure (contact) distributions with the probability that an exposure to a given compound occurs as a result of a previous or concurrent application of a product containing that chemical. The model also can take into account the probability that exposures to more than one product may occur on a single day, which provides a more realistic exposure estimate than if exposures from single uses are

Hayes’ Handbook of Pesticide Toxicology

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Residue (per unit of contact)

Treatment 1

Treatment 2 Treatment 3 R3 R1 R2

1

90

100

180 Days

250 270

365

Day 250 - Residue = R2 + R3

Figure 50.4  Available residue for one application scenario of three treatment types with chemical X in 1 year.

summed. Monte Carlo techniques are used to estimate the distribution of potential exposures (contacts) and to combine these distributions with information about product use. These conditional distributions are combined with the usage probabilities associated with each product to generate exposure distributions for specific calendar days specified by the user and are repeated many thousands of times in the analysis.

50.2.11  Toxicity Data Used to Place the Exposure Assessment in Context Ideally, exposure estimates are compared to compoundspecific toxicity measures to derive risk estimates. The most appropriate toxicity measure depends on the type of assessment being conducted. Examples of toxicity values are dermal, oral, and inhalation no observed effect levels (NOELs), effective doses (ED50), or reference doses (RfDs). Different risk estimates may be needed to place the compounds in proper perspective for either aggregate or cumulative assessments compared to when a compound is evaluated individually. Estimates of chronic exposure are typically compared to a chronic reference dose and may be expressed as a percentage of that reference dose. Estimates of acute exposures are usually compared to the acute NOEL. Acute risk estimates are expressed as margin of exposures (MOEs). Estimates of short- and intermediate-term exposures are usually compared to the short- and intermediate-term NOELs, respectively.

50.3  Sample model calculation methodologies To provide better guidance, this section presents a representative set of the methodologies and algorithms. The analyst will want to select the exposure methodology that is most appropriate for the desired analysis.

50.3.1  Typical Exposure Methodology Sequence Prior to the analysis: 1. Define product- and route-specific parameters for each source of the pesticide. Each exposure route (dermal, oral, or inhalation) of concern for each treatment type (e.g., treatment of agricultural crops, pet treatment, turf treatment, crack and crevice treatment) must be identified. If the scenario represents different exposure routes for the same treatment type, the model should have the same set of application scenarios, but will likely utilize different data. 2. Determine all treatments of interest (e.g., crops that are treated for dietary assessment and all nondietary uses of the chemical for aggregate assessment). 3. Select the most appropriate software to use to determine exposure. 4. Identify available contact and residue data to use in the assessment (see Figure 50.5). Conduct the analysis: 5. Define analysis parameters. Typical parameters to be defined are the type of analysis of interest (given in the following sublist), the number of Monte Carlo iterations and random number seed, the appropriate NOEL or RfD values by exposure route, and the subpopulations of interest (e.g., toddlers 1–3, southern United States). Typical analysis types include: • Single day. In this analysis, exposure is calculated for a randomly chosen day where all days of the year are given equal probability. Exposure also may be calculated for specific sequential days if that is appropriate. Typically, the output provides a distribution of the daily exposures for each of the specified days. • Multiple weeks. Exposure is calculated for a specified week or combined set of weeks (up to 52 combined weeks) that are chosen by the user. The

Chapter | 50  Modeling Dietary Exposure with Special Sections on Modeling Aggregate and Cumulative Exposure

CSFII Demographics

Day (1),

CSFII Food Consumption

Day (2),.....

Day (ND)

Dietary

DEEM Residue File Treatment (1) Contact Library Residue Library

AGX (1)

Inhalation

AGX (2)

Dermal

AGXx (3)

Oral

Treatment (2) Contact Library Residue Library

AGX (4)

Inhalation

AGX (5)

Dermal

AGX (6)

Oral

Treatment (NT) Contact Library Residue Library

AGX (N-2)

Inhalation

AGX (N-1)

Dermal

AGX (N)

Oral

Summarize and Bin Results

Figure 50.5  Compute exposure.

output can provide a distribution of the daily exposures averaged over the combined number of days. Exposure also can be calculated for a series of weeks or a series of a combined set of weeks.   6. Dietary exposure is usually an important source of exposure, and the analyst may conduct the dietary exposure separately prior to conducting aggregate exposure assessments.   7. Exposure calculations are usually repeated for each treatment or exposure, and the total exposure is calculated and stored for later use to estimate a distribution and to estimate exposure by route (dermal, inhalation, oral). For all analyses:   8. Determine the average daily exposure (by route). One approach is to calculate the average daily exposure by dividing the sums of the multiple-day exposures by the number of days (n) specified in step 3.   9. Express exposure on a unit body weight (BW) basis. 10. For cumulative exposures, determine how relative potency will be expressed and adjust residues to reflect the differences.

50.3.2  Monte Carlo Simulation Methodology Monte Carlo analysis methods are used to bring together the wide range of probability distributions needed to calculate each individual’s exposure. Specifically, for each

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treatment type (e.g., agricultural crop treatments, residential turf treatments) included in the analysis, Monte Carlo analysis can be used to define the use scenarios. For example, Monte Carlo sampling could be used to decide (1) whether the pesticide is applied to each crop, (2) whether it is applied professionally or by someone in the home, (3) whether that treatment type is made at the individual’s residence, (4) how many times a year the treatment is made, (5) the dates that each treatment is made during the year, (6) what residue amount results from each treatment on the day(s) of treatment, and (7) how much contact with potentially residue-bearing surfaces (or other residue transfer media) is made by the individual on the days of interest. Ideally, the analyst determines whether each of these decision variables is to be based on a deterministic value or on a distribution from which a value is drawn using a new random number with each draw. Often food consumption amounts used in dietary analyses are treated as deterministic values, whereas the residues assigned to those food consumption amounts can be either deterministic or drawn from a distribution. In addition, most, if not all, demographic parameters associated with each respondent are considered to be deterministic. That is, the person’s reported weight and height are used in the analysis, not simulated using a probabilistic distribution.

50.3.3  Degradation Calculation Methodology The degradation calculation model incorporates residue degradation by using either actual residue values on specific days after treatment or degradation equations. If market basket data are available, degradation equations are not required to estimate dietary exposure. Degradation equations are important for most nondietary evaluations. The degradation equations are as follows: For the half-life method,

RX  R0 (0.5 X/hl )

(2)

where hl denotes half-life in days. For the straight-line method,

RX  R0 (1  X / z )

(3)

where z represents days to zero residue level. RX is determined directly from a residue distribution function other than R0: Given that ldi is the last day that residue function i is valid, find Ri such that ld(i – 1)  X  ld(i). Then RX  Ri (selected from probability distribution), where X is the integer number of days elapsed between application and contact, R0 is the residue concentration factor on day of application (mg/unit of contact), and RX is the residue concentration factor on day X (mg/unit of contact).

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Note: When X  0, RX  R0 with no adjustment. That is, on the day of application, the treatment was assumed to be applied before contact and the residue is not degraded.

50.3.4  Risk (MOE) Calculation Methodology When a single exposure route is evaluated for a chemical or multiple routes are evaluated in which the NOEL for each route is the same, the MOE can be calculated for any exposure value in the resulting distribution. For example, if the 90th percentile exposure amount is Exp90, then the MOE is simply NOEL/Exp90. However, when multiple exposure routes that each have a different NOEL are evaluated, the total MOE cannot be calculated directly for any given exposure in the exposure distribution. Instead, the MOE must be calculated for each individual from the individual components of his or her total exposure, that is,

50.4.2  Parameter Uncertainty Parameter uncertainty includes measurement errors, sampling errors, variability, and use of surrogate data. Two examples of measurement uncertainty in the data may be the presumed tendency of some survey respondents to underestimate their body weights or to underreport food consumption. In the first example, parameter uncertainty may result in potential overestimation of the exposures, whereas in the second example, it may result in potential underestimation of exposures. Sampling errors may result from sampling too few observations or nonrepresentative sampling. Generally, studies of residential exposures often include very few measurements and typically are conducted for a limited number of scenarios.

50.4.3  Model Uncertainty

A comparison of the results of the various analyses provides MOE1  Exp1 /NOEL1 , MOE 2  Exp2 /NOEL 2 , MOE 1  Exp3 /NOEL 3 the assessor with a measure of the impact of the uncer tainty in the exposure model used. Another example of where subscript 1 refers to the inhalation exposure route, model uncertainty may result from using the wrong model 2 to the dermal exposure route, and 3 to the oral exposure to represent the degradation, over time, in air (or soil) route (which includes both dietary and incidental inges- concentrations. tion). Then the total MOE (MOET) can be calculated based on the equation in the Office of Pesticide Programs draft 50.5  Example of aggregate exposure Guidance for Performing Aggregate Exposure and Risk Assessments (U.S. EPA, 1999b), which is 50.5.1  Description



 1 1 1   MOE T  1 /     MOE1 MOE 2 MOE n 



(4)

where, for example, MOE1 denotes a given margin of exposure (e.g., adult, inhalation route) and MOE2 denotes another specified margin of exposure (e.g., adult, dermal route). This total MOE (MOET) concept was presented to and endorsed by FIFRA’s SAP in 1997 (U.S. EPA, 1999b).

50.4  Uncertainty The EPA (1992) has classified uncertainty in exposure assessments in three categories: scenario uncertainty, parameter uncertainty, and model uncertainty. Examples of how these uncertainties may arise in an exposure assessment follow.

50.4.1  Scenario Uncertainty Scenario uncertainties include descriptive errors, aggregation errors, and incomplete analysis. For instance, for residues on imported crops, scenario uncertainty may result from incorrect information regarding the regions in which the product is used and how it is used.

This example illustrates an approach that is used to aggregate exposures for adults (18 years old) and toddlers (1–3 years old) exposed to residues of chemical X from a hypothetical turf product that is also applied to fruit crops. The product is a granular formulation used to control weeds and is applied using a push-type spreader. It was assumed that the homeowner makes five applications of this product at 4-week intervals, with the first application in the first week of May. Therefore, the potential exposures that result from the use of this product include dermal and inhalation adult applicator exposures, dermal postapplication exposure for adults and toddlers, and toddler incidental ingestion via hand-to-mouth behavior. For simplicity, we did not include the other oral exposures as specified in the EPA residential (SOPs; U.S. EPA, 1999a) for the toddlers. Chemical X was assumed to be completely degraded after 35 days. It was also assumed that the toxic effects from the dermal, oral, and inhalation routes are the same; therefore, the exposures from these routes were aggregated together. The dermal NOEL was assumed to be 85 mg/kg BW/day, whereas the inhalation and oral NOELs were 20 mg/kg BW/day. For simplicity it was assumed that the short- and intermediate-term NOELs are the same. Calendex software, developed by Durango Software and licensed by Exponent, Inc., was used to calculate exposure. The exposures were estimated for a single randomly chosen day

Chapter | 50  Modeling Dietary Exposure with Special Sections on Modeling Aggregate and Cumulative Exposure

(Table 50.1) and for five 4-week periods (weeks 18–21; weeks 22–25; weeks 26–29; weeks 30–33; and weeks 34– 37; Table 50.2). The EPA SOP (EPA, 1997a) equations and parameters, which are presented subsequently, were used as the basis for the nondietary exposure algorithms. The distribution types and sources of data are listed below each parameter.

50.5.2  Adult Applicator Dermal and Inhalation Exposures UE(dermal or inhalation)  AR PDR  (5) BW



where PDR  potential dose rate (mg/kg per day); UE  unit exposure (dermal or inhalation) (mg/lb ai), point estimate: Pesticide Handlers Exposure Database (Versar, 1995), UEdermal  3 mg/lb ai, UEinhalation  6.3 mg/ lb ai; AR  application rate (lb ai/acre), hypothetical uniform distribution: 0.5–1.5 lb ai/acre; A  area treated (acre/day),w hypothetical uniform distribution: 0.25–1.0

acre/day; and BW  body weight (kg), empirical distribution: CSFII data.

50.5.3  Adult and Toddler Postapplication Dermal Exposures PDR 



(DFR  TC  ET  CFI) BW

Adults Exposure

Toddlers MOE

Percentile

Exposure

MOE

Dermal (postapplication)

Dermal (applicator  postapplication) 99.9

0.459265

185

99.9

0.795279

107

99

0.316427

269

99

0.539392

158

95

0.199563

426

95

0.312677

272

90

0.140733

604

90

0.205951

413

99.9

0.090714

220

99.9

0.029207

685

99

0.000000

N/A

99

0.020774

963

Inhalation (applicator)

Oral (nondietary)

95

0.000000

N/A

95

0.011949

1674

90

0.000000

N/A

90

0.008066

2479

0.046856

427

Oral (dietary)

Oral (dietary) 99.9

0.004861

4114

99.9

99

0.001313

15,238

99

0.018387

1088

95

0.000182

109,690

95

0.005180

3861

90

0.000029

681,860

90

0.001630

12,266

Aggregate (dermal  inhalation  oral)

(6a)

where PDR  potential dose rate (mg/kg per day); DFR  dislodgeable foliar residue (g/cm2) [see Eq. (50.6b)]; TC  transfer coefficient (cm2/h), point estimates: EPA SOPs, adult short term  14,500, adult intermediate term  7300, toddler short term  5200, toddler intermediate term  2600; ET  exposure time (h/day), cumulative distribution of amount of time spent playing on grass shown in table below, from EPA Exposure Factors Handbook (EPA, 1997b); CF1  conversion factor (0.001 mg/g); and BW  body weight (kg), empirical distribution: CSFII data.

Table 50.1  Results of the Turf and Dietary Aggregate Example: Single-Day (Randomly Selected) Exposures (mg/kg BW/day; per capita) Percentile

1111

Aggregate (dermal  oral)

99.9

190

99.9

100

99

280

99

150

95

440

95

250

90

620

90

360

Hayes’ Handbook of Pesticide Toxicology

1112

Adults

Toddlers

Percentile

ET

Percentile

ET

0

0

0

0

25

0.5

25

0.25

50

2.0

50

1.0

100

2.0

75

2.0

100

2.0

DFR  AR  F  CF 2  CF3



(6b)

where DFR  dislodgeable foliar residue (g/cm2); AR  application rate (lb ai/acre), hypothetical uniform distribution: 0.5–1.5 lb ai/acre; F  fraction of ai transferred from foliage (%), point estimate: EPA SOPs, 5%; CF2  conversion factor (4.54  108 g/lb); and CF3  conversion factor (24.7  109 acre/cm2).

50.5.4  Toddler Postapplication Incidental Ingestion Exposures PDR 

(DFR  SA  FQ  ET  SE  CF1) BW

(7)

where PDR  potential dose rate (mg/kg per day); DFR  dislodgeable foliar residue (g/cm2) [see Eq. (6b)]; SA  surface area of two fingers (cm2/event), point estimate: EPA SOPs, 20 cm2; FQ  frequency of hand-to-mouth activity (events/h), point estimate: EPA SOPs, 20 events/h; ET  exposure time (h/day), cumulative distribution of amount of time spent playing on grass shown in table below, from Exposure Factors Handbook; SE  saliva extraction (%), point estimate: EPA SOPs, 50%; CF1  conversion factor (0.001 mg/g); and BW  body weight (kg), empirical distribution: CSFII data.

Toddlers Percentile

ET

0

0

25

0.25

50

1.0

75

2.0

100

2.0

50.5.5  Reporting Results A comprehensive report of each exposure analysis should always be prepared to fully document the assumptions and data that were used and to link that information to

the actual analysis. The report should include estimates of exposure and risk (presented as an MOE, %NOEL, %RfD, or other toxicity parameter) for each subpopulation identified by the analyst. The exposure and risk estimates should be presented for the full range of the population. For an individual treatment (e.g., treatment of apples or lawn treatment or only pet treatment), exposure and risk estimates are presented for each individual route (e.g., dermal, inhalation, incidental ingestion) as well as any aggregated routes (e.g., inhalation and incidental ingestion combined). For an aggregate assessment (e.g., lawn treatment  pet treatment  dietary), exposure and risk estimates should be presented for the individual routes, but the estimates are aggregated by uses (e.g., the dermal estimate would be an aggregation of the dermal from lawn care plus the dermal from pet treatment). Additionally, if there were similar toxicological endpoints from both dermal and oral exposure, then these estimates could be combined as well. In that case, there could be a total aggregation.

50.6  Quality Audit and Validation Audits of the computational algorithms used in an assessment must be derived using an independent spreadsheet calculation or other software. These include the algorithms for deriving the interval limits, allocation of the observations to the intervals, and calculation of the various statistics, including the means, standard deviations, and percentile estimates.

50.7  Commercial and Government Exposure Assessment Software The analyses described previously are data-intensive and require thousands of simulations that are best accomplished using a computer. Several custom computer tools are available that have been designed specifically for estimating exposures to pesticides. The EPA has conducted an extensive review of the software summarized in the following and concluded that each of these tools can be used to estimate exposure for submission to the EPA. Two models focus entirely on dietary exposure (DEEMFCID and DEPM). The other models (Calendex, CARES, REXTMLifeLine, and EPA SHEDS-Wood) include tools that allow the assessor to estimate dietary, aggregate and cumulative exposures. Each of the models is described briefly in the following sections. The reader can find more information on the website for each of the models. The models are presented in alphabetical order.

50.7.1  Calendex Durango Software, LLC expanded its analytical software capabilities with the introduction of Calendex in order

Chapter | 50  Modeling Dietary Exposure with Special Sections on Modeling Aggregate and Cumulative Exposure

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Table 50.2  Results of the Turf and Dietary Aggregate Example: 4-Week Exposures (mg/kg BW/day; per capita) Adults Percentile

Exposure

Toddlers MOE

Percentile

343

99.9

0.247649

MOE

Dermal (postapplication)

Dermal (applicator  postapplication) 99.9

Exposure 0.386618

220

99

0.206313

412

99

0.310551

274

95

0.170271

499

95

0.256527

331

3339

99.9

0.015054

1329

Inhalation (applicator) 99.9

0.005990

Oral (nondietary)

99

0.004656

4296

99

0.012200

1639

95

0.003526

5673

95

0.009980

2004

99.9

0.001410

14,185

99.9

0.012927

1547

99

0.000649

30,801

99

0.007424

2694

95

0.000273

73,202

95

0.003905

5121

Oral (dietary)

Oral (dietary)

Aggregate (dermal  inhalation  oral)

Aggregate (dermal  oral)

99.9

310

99.9

180

99

375

99

226

95

457

95

276

to provide tools to meet the requirements of the EPA for conducting aggregate and cumulative exposure analyses. Calendex is a calendar-based software package to use to estimate aggregate and cumulative exposure analyses of dietary and residential chemicals. Together, Calendex and DEEM-FCID offer the risk assessor a set of analytical tools for exposure analysis. The programs allow extensive documentation of the user’s input data, flexible exposure modeling tools, a wide range of computation options, and clear, concise reports. DEEM and Calendex have been made publicly available and use only publicly available databases (such as the USDA CSFII and EPA’s FCID). Additional information about Calendex is available at http://www.durango-software.com/software/calendex.html.

50.7.2  Cumulative and Aggregate Residue Evaluation System (CARES) CARES is a software program designed to conduct complex exposure and risk assessments for pesticides, such as the assessments required under the 1996 Food Quality Protection Act (FQPA). CARES was originally developed under the auspices of CropLife America (CLA), which conceived the project, provided funding, and oversaw the program’s evolution. CropLife America has made the CARES source code freely available to all stakeholders and government agencies. There are several features that differentiate CARES

from other risk assessment models. The Notitia, a proprietary product of infoscientific.com, Inc., serves as the software engine to run CARES and to house data. The Notitia framework provides the flexibility to introduce and link exposure scenarios to assess their potential impact for products already in the marketplace as well as for new registration candidates still in evaluation. The population generator, POP GENTM, from Sielken & Associates Consulting, Inc., is based on the U.S. Census and enables CARES to match individual and population attributes across single or multiple exposure databases so that the user can generate a 1-day or 365-day profile. The software provides tools to estimate the contribution of different exposure sources as well as an analysis module to evauate the sensitivity of assumptions. The CARES program and source code is to be publicly available at no charge. CARES can be downloaded at http://cares.ilsi.org/.

50.7.3  Dietary Exposure Evaluation Model (DEEM)-FCID DEEM-FCID s a dietary exposure analysis system for performing chronic and acute exposure assessments. DEEM employs Monte Carlo Analysis (MCA) techniques in order to provide probabilistic assessments of dietary pesticide exposures when residue data for targeted foods are available as distributions.

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DEEM can also perform cumulative exposure analyses when the contribution of multiple chemicals used on different foods to the total exposure must be evaluated. The platform was developed by Durango Software, LLC. DEEM is a user-oriented, menu-driven system that runs on IBM personal computers and compatible microcomputers. It operates under virtually all Windows operating systems. DEEM incorporates the 1994–1996, 1998 CSFII with USDA-EPA FCID recipes to translate the foods as eaten to RACs and foodforms. DEEM-FCID consists of four software modules: the main DEEM-FCID module, the acute analysis module, the chronic analysis module, and the RDFdoc utility for validating and documenting residue distribution files (RDFs). The main DEEM-FCID module is used to create and edit residue files for specific chemical or cumulative applications and to launch the DEEM acute, chronic, and RDFdoc modules. In addition, a related program, the RDFgen residue utility, can be used to automate single analyte and cumulative residue distribution adjustments and the creation of summary statistics and RDFs based on USDA Pesticide Data Program (PDP) monitoring data or user-provided residue data. The DEEM software itself is also used with Calendex, a cumulative and aggregate exposure assessment software application focusing on combined dietary and (nondietary) residential exposures. Additional information about DEEM-FCID may be found at http://www.durango-software.com/software/deem.html.

50.7.4  Dietary Exposure Potential Model (DEPM) DEPM is a model using extant food databases to estimate dietary exposure to chemical residues that can be used for identifying the importance of diet relative to other exposure pathways and indicating the potential for high exposure of certain populations. Existing consumption and contaminant residue databases, normally developed for purposes such as nutrition and regulatory monitoring, contain information to characterize dietary intake of environmental chemicals. A model and database system, termed the Dietary Exposure Potential Model (DEPM), correlates extant food information in a format for estimating dietary exposure. The resident database system includes several national, government-sponsored food intake surveys and chemical residue data from monitoring programs. A special feature of the DEPM is the use of recipes developed specifically for exposure analysis that link consumption survey data for prepared foods to the chemical residue information, which is normally reported for raw food ingredients. Consumption in the model is based on 11 food groups containing approximately 820 exposure core food (ECF) items with similar basic ingredients, established from mean values of consumption of over 6700 food items commonly identified in food surveys. The summary

Hayes’ Handbook of Pesticide Toxicology

ECF databases are aggregated in a fashion to allow analyst selection of demographic factors, such as age/gender groups, geographical regions, ethnic groups, and economic status. Daily intake is estimated by the model based on mean values of residues reported for over 350 pesticides and environmental contaminants. In addition, contributions to estimated dietary exposure from ECF groups and individual ECFs can also be estimated. Though not intended for risk analysis, the model has proved to be a suitable tool for designing and interpreting exposure measurements, identifying data gaps, and establishing priorities for dietary exposure research.

50.7.5  Lifeline The LifeLine Group has created software tools for a variety of exposure assessment applications. Information on the LifeLine Group webware includes a description of new software tools that allow updating of dietary surveys which incorporate data for unique or atypical populations, and for conducting dietary exposure estimation and dietary risk assessment for these diets as well as for blended diets. A stand-alone software tool, called the Dietary Record Generator (DRG), can convert available dietary information into comprehensive dietary intake profiles. The resulting e databases that are generated by the DRG can be used as if they were generated from a conventional survey such as the U.S. Department of Agriculture’s Continuing Survey of Food Intakes for Individuals (CSFII). The LifeLine Group has developed the LifeLine Customized Dietary Assessment Software (CDAS). This software allows the user to calculate dietary exposure and risks from unique diets. Another important feature of the CDAS is that it is able to blend the new dietary information from the DRG database with the dietary information contained in the CSFII database and evaluate the dietary exposure and risk of the chemicals that are contained in the blended diet. Using the DRG software, excellent food consumption databases have been constructed from a diverse array of information about dietary habits, food availability, and economics of the populations for whom there are no detailed food consumption surveys. These profiles and the basis upon which they were built are compiled in a comprehensive document entitled The Compendium of Alaska Traditional and Subsistence Dietary Files. The DRG and the CDAS address two issues of major concern in dietary risk assessment. The DRG provides a cost-effective mechanism for updating existing survey databases to reflect changes in food consumption, as well as being able to generate food consumption databases for many groups, including other ethnic and cultural groups or people on unique or atypical diets such as vegetarian diets, ethnic diets, low-carbohydrate diets, diets high in any specific food of interest, or high in fortified foods. The CDAS

Chapter | 50  Modeling Dietary Exposure with Special Sections on Modeling Aggregate and Cumulative Exposure

can then be used to assess dietary exposure and risks in any of these diets. Additional information about these programs can be found at http://www.thelifelinegroup.org/newapproach.htm.

1115

be obtained from the U.S. EPA website: http://www.epa. gov/heasd/risk/projects_completed/sheds_wood_cca.htm.

Summary 50.7.6  REx (Residential Exposure Assessment Model) REx is a spreadsheet (Excel)-based exposure/dose assessment model that allows aggregating product use scenarios. REx allows deterministic and complex stochastic exposure and dose assessments to be conducted. REx is based on EPA’s Residential SOPs. One or more (up to six) scenarios can be aggregated to estimate exposure and dose to receptors of interest. Subgroups available in in REx are adults and three categories of children (age  1 year, 1 year  age  6 years, and age  6 years). Aggregation can be done for day 0 during and post application scenarios. Additional information about Rex can be found at http://www.infoscientific.com/software_main.htm.

50.7.7  U.S. EPA Sheds-Wood Model The SHEDS-Wood model is a probabilistic model designed to simulate aggregate exposures and doses for population cohorts and multimedia chemicals, by using data from time-location-activity diaries compiled in the U.S. EPA’s Consolidated Human Activity Database (CHAD) (McCurdy et al., 2000). The EPA’s Office of Research and Development (ORD), National Exposure Research Laboratory (ORD/NERL) developed and applied the probabilistic Stochastic Human Exposure and Dose Simulation model for wood preservatives (SHEDS-Wood) to estimate children’s absorbed dose of the arsenic and chromium components of CCA. Skin contact with, and nondietary ingestion of, arsenic and chromium in soil and wood residues were considered for the population of children in the United States who frequently contact chromated copper arsenate (CCA)-treated wood playsets and decks. Model analyses were conducted to assess the range and uncertainty in population estimates, key model inputs, and the impact of various potential mitigation strategies such as the use of sealants and hand washing after play events. The model input parameters are grouped in four categories: (a) activity factors (including parameters such as the average number of days that a child plays on or near playsets/ decks, fraction of time a child is actually on the playset/ deck), (b) exposure factors (including frequency of handto-mouth activity, washing events, dermal loading, residue skin transfer efficiency), (c) dose factors (including dermal and gastrointestinal absorption fractions), and (d) environmental media parameters (including soil and residue concentrations). Additional information about the model can

Consumer exposures to pesticides can occur as a result of residues in food, water, and the air, as well as residues on surfaces in the home and garden. These exposures can be estimated using models which have been developed specifically for this purpose. The models allow “worst case” estimates as well as refined “most realistic” estimates of exposure. Consumer exposures can be estimated for a single day’s exposure as well as weekly, monthly or even a lifetime. The most appropriate estimate will depend upon the toxicological issue to which the exposure is to be compared as well as the intended application (e.g. for regulatory compliance). Typically the more refined the estimate, the more data that are required to conduct the evaluation. Regardless of the model selected for an analysis, the selection of the data and related assumptions will determine the quality of the assessment. In all cases, uncertainty and variability should be described. Assumptions should be documented. The results should be compared to appropriate estimates of the toxicity of the pesticide.

References Australia New Zealand Food Authority (ANZFA). (1997). “Dietary Modelling: Principles and Procedures”. Australia New Zealand Food Authority, Canberra, Australia. Food Standard Australia New Zealand, 2008. NUTTAB 2006 Electronic Release. Available at: http://www. foodstandards.gov.au/monitoringandsurveillance/nuttab2006/electronicrelease.cfm. Australia New Zealand Food Authority (ANZFA). (2009). Website describes the food consumption surveys: http://www.foodstandards.gov.au/ monitoringandsurveillance/dietaryexposureasses254.cfm. Bassett, M. T., Dumanovsky, T., Huang, C., Silver, L. D., Young, C., Nonas, C., Matte, T. D., Chideya, S., and Friede T. R. (2008). In “Purchasing Behavior and Calorie Information at Fast-Food Chains in New York City.” American Journal of Public Health 98,1457–1459. China (2002). Chinese Ministry of Health. China National Nutrition and Health Survey, Beijing, China. Data available at: http://www.cpc.unc. edu/projects/china. China (2006). Chinese Ministry of Health. China National Nutrition and Health Survey. Beijing, 2006. Data available at: http://www.cpc.unc. edu/projects/china. Data Food Networking II (DAFNE) (1995). “Network for the Pan-European Food Data Bank Based on Household Budget Surveys.” At: 12http:// ec.europa.eu/health/ph_projects/1999/monitoring/fp_monitoring_1999_ annexe_uk_01_en.pdf. FAO STAT from 1960-2008 FAO Balance Sheet data (FAOSTAT.PC), 2007, Various versions are available through the FAO website FAO: stats general are available http://apps.fao.org.

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Federation of American Societies of Experimental Biology (FASEB)(1988). “Estimation of Exposure to Substances in the Food Supply.” Life Sciences Research Office, Bethesda, MD. Japan (2005). Ministry of Health, Labour and Welfare of Japan (2005), The National Health and Nutrition Survey in Japan, 2005, Tokyo, Japan. Lin, B., Guthrie, J., and Frazao, E. (1999). Nutrient contribution of food away from home. In “America’s Eating Habits: Changes and Consequences,” Frazao, ed.), pp. 213–242. Washington, DC, US Dept of Agriculture. Ministry of Health, Labour and Welfare of Japan. (2005). The National Health and Nutrition Survey in Japan, 2005, Tokyo, Japan. Mosteller, F., and Tukey, J. W.(1977). “Data Analysis and Regression,”. Addison–Wesley, Reading, MA. National Center for www.cdc.gov/nchs/nhanes.htm. Pao, E. M., Sykes, K. E., and Cypel, Y. S. (1989). “USDA Methodological Research for Large-Scale Dietary Intake Surveys. 1975–88.” Home Economics Research Report No. 49. U.S. Department of Agriculture, Human Nutritition Information Service, Washington, DC. Samet, S. M. (1989). Surrogate measures of dietary intake. Am. J. Clin. Nutr. 50, 1139. Sasaki, S., and Kestelhoot, H. (1992). Value of Food and Agriculture Organization data on food-balance sheets as a data source for dietary fat intake in epidemiologic studies. Am. J. Clin. Nutr., 56, 716. Trichopoulou, A., and Pagona, L. (eds.) (1997). “Methodology for the Exploitation of HBS Food Data and Results on Food Availability in 5 European Countries.” DAFNE (Data Food Networking) European Communities. United States Department of Agriculture (USDA) (1992). “Nationwide Food Consumption Survey: Continuing Survey of Food Intakes by Individuals 1989–90.” Dataset, Human Nutrition Information Service. USDA (1993). “Nationwide Food Consumption Survey: Continuing Survey of Food Intakes by Individuals 1990–91.” Dataset, Human Nutrition Information Service. USDA (1994). “Nationwide Food Consumption Survey: Continuing Survey of Food Intakes by Individuals 1991–92.” Dataset, Human Nutrition Information Service.

Hayes’ Handbook of Pesticide Toxicology

USDA (1995). “Nationwide Food Consumption Survey: Continuing Survey of Food Intakes by Individuals 1993–94.” Dataset, Food Survey Research Group, Agricultural Research Service. USDA (1996–1998). “Nationwide Food Consumption Survey: Continuing Survey of Food Intakes by Individuals 1994–95.” Dataset, Food Survey Research Group, Agricultural Research Service. U.S. EPA (United States Environmental Protection Agency) (1997a). “Standard Operating Procedures for Residential Exposure Assessments.” Residential Exposure Assessment Work Group, Office of Pesticide Programs, Health Effects Division, EPA, Washington, DC. (Draft version dated December 18, 1997.) U.S. EPA (2009). The Sheds Woods Program. http://www.epa.gov/heasd/ risk/projects_completed/sheds_wood_cca.htm. U.S. EPA (1997b). “EPA Exposure Factors Handbook.” Office of Research and Development, EPA, Washington, DC. U.S. EPA (1999a). “Standard Operating Procedures for Residential Exposure Assessments – Proposed Changes Presented to Science Advisory Panel.” Office of Pesticide Programs, Health Effects Division, EPA, Washington, DC. U.S. EPA (1999b). “Guidance for Performing Aggregate Exposure and Risk Assessment.” Office of Pesticide Programs, EPA, Washington, DC. U.S. Food and Drug Administration. (2006). Estimating Dietary Intake of Substances in Food. Available at: http://www.cfsan.fda.gov/~dms/ opa2cg8.html. Versar (1995). “PHED: The Pesticide Handlers Exposure Database,” Reference Manual, Version 1.1. Prepared for the PHED Task Force that represented Health Canada, USEPA, and American Crop Protection Association. WHO Exposure Assessment Guidelines. (2008). Report of workshop which will form the basis for the chapter on dietary exposure assessment for the guidance document on ‘Principles and Methods on the Risk Assessment of Chemicals in Food’, which is currently being finalized and hopefully will be published late this year. The report can be found at: http:// whqlibdoc.who.int/publications/2008/9789241597470_eng.pdf.

Chapter 51

Modern Approaches to Analysis of Pesticide Residues in Foods and the Environment Luis O. Ruzo PTRL West, Inc. Thomas Class PTRL Europe

51.1  Introduction The capability to determine the actual residues of pesticides, and increasingly of their metabolites, is at the core of the complex and sometimes bewildering regulatory processes in all industrialized nations. Thus, in the United States we are currently experiencing the replacement of laws based on long-term toxicological effects, such as the Delaney Clause, with those based on quantitative determinations that lead to a different type of risk assessment, as is the Food Quality Protection Act of 1996 (FQPA). To understand the benefits and limitations of modern analytical techniques, it is necessary to examine the regulatory requirements that they aim to satisfy. Residue chemistry data are used by regulatory agencies to estimate the exposure of the general population, as well as discrete subpopulations, to pesticide residues in food and water and for setting and enforcing tolerances for such residues in food crops or animal feed. These data are also used to monitor environmental contamination of soil, air, and water and thus to determine adverse effects that may arise from transport of residues between these compartments. The tolerance values are a key component of the regulatory equation as they represent the amounts of pesticide-related materials legally allowed to be present in a given matrix. Thus, tolerances are legally enforceable limits that are currently under scrutiny for implementation of the FQPA. The passage of this law by the U.S. Congress in 1996 signaled a fundamental change in the way exposure to pesticides is evaluated by introducing the concept of aggregate exposure to several compounds of a given chemical class and mode of action. These expanded requirements will necessitate the development of new methodologies for analysis, specifically to address the lower limits of quantitation (LOQs) (Ragsdale, 1998). Hayes’ Handbook of Pesticide Toxicology Copyright © 2001 Elsevier Inc. All rights reserved

A similar situation is observed in the European Community (EC). Here (re-)registration of pesticides and enforcement of tolerances or maximum residue limits (MRLs) requires that existing multiresidue methods be assessed for their applicability toward the determination of active substances and their toxicologically relevant metabolites.

51.2  Method validation Because this chapter will deal primarily with analytical techniques now in use for developing methodologies, it is worth reviewing briefly the criteria that enforcement agencies use to evaluate method performance. Methods used by government laboratories are generally developed by the pesticide registrants. A method is considered acceptable upon validation. This process may entail the examination of various parameters but always includes the establishment of a limit of quantitation (LOQ) and a limit of detection (LOD). The former is usually set at the lowest matrix fortification level for which acceptable, quantifiable recoveries of the analyte(s) are obtained.

51.2.1  Confirmation of the Analytical Method This entails the fortification of untreated (control) matrices (crops, soil, water, etc.) with varying concentrations of analyte. Thus, the processed sample is fortified at two or three levels in duplicate or triplicate (United States) or in five replicates (EC). The fortified samples and corresponding untreated controls are then subjected to extraction, cleanup, and chromatographic separation and quantitation. 1117

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The ultimate recovery of analytes must be in the 70–120% (70–110% in Europe) range based on the initial concentration, with repeatability demonstrated by relative standard deviations of  20%. The lowest concentration for which this is achieved during validation is generally considered the limit of quantitation or determination (LOQ) for the method. The limit of detection (LOD) may be any value below the LOQ at which an analyte signal is clearly distinguished from background signals present or absent in matrix extracts. At present, methods aim at an LOQ of 0.01 ppm (mg/kg) for most foodstuffs, whereas much lower values are considered desirable for water analyses, where the LOQ  0.05 g/1 (ppb) is required, or for compounds of known toxicological significance. Important aspects to be considered in the conduct of validation studies are provided by Jenke (1996a–c).

51.2.2  Independent Laboratory Validation For a method to be used in the enforcement of tolerances or MRLs, it must be rugged and straightforward. Therefore, the U.S. Environmental Protection Agency (EPA) and the EC require that such methods be validated independently by laboratories that are not familiar with the procedures or analytes (EC Directorate General for Agriculture, 1998; EPA, 1996). Because the equipment necessary for the enforcement method must be generally accessible and affordable for the enforcement laboratories, some of the most advanced techniques (such as MS/MS, or in the European Community LC-MS) are still not acceptable to regulatory agencies. Measures to demonstrate the validity of results obtained by enforcement methods include round-robin testing using identical samples with fortified or incurred residues, which allow assessment of the reproducibility of commonly employed enforcement methods.

51.2.3  Method Radiovalidation (EPA, 1996) A stringent test for an analytical method is its reproducibility when applied to incurred residues as opposed to an externally fortified matrix. Plant and animal metabolism studies utilizing 14C-labeled pesticides generate matrices containing incurred residues that can be readily quantified with radiochemical methodology, which is quite different (and simpler) than that generally developed for an analytical residue method. In order for a method to be considered fully validated, the results obtained when the “cold” method is applied to matrices containing 14C-labeled incurred residues must agree closely with the results arising from quantitation of the radiocarbon conducted by radiochemical methods such as liquid chromatography and liquid scintillation counting.

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51.3  Development of the analytical method 51.3.1  Multiresidue Methods Multiresidue, multiclass methods are generally the most cost effective overall approach for pesticide analysis in foods, soil, and water. Regulatory enforcement methods routinely deal with multiresidue determinations under fairly standardized conditions. For example, the widely used DFG S19 method (Thier and Zeumer, 1987) establishes conditions for extraction and quantitation of organochlorine, organophosphorous, and nitrogen-containing pesticides (typically 80–100 compounds per method) in crop plants under standardized conditions. The original method (Specht et al., 1995) uses an acetone/ethyl acetate/cyc-lohexane mixture for extraction and partition, thus replacing dichloromethane and introducing a “one-beaker” extraction and partition procedure. In addition to use with watery plant matrices, the DFG S19 method has also been successfully employed for dry plant matrices (straw, hops, tobacco, herbal teas), for oily crops (oilseed rape, sunflower seeds, nuts, etc.), for animal matrices (milk, whole egg, muscle, fat), and for soil. Thus, a single solvent (acetone) with fixed amounts of added water (in a ratio of 2:1 solvent:water) or acetonitrile (for oily matrices) is used for extraction of compounds with a wide range of polarity. Cleanup involves gel permeation chromatography (GPC) and fractionation on a small silica gel column followed by separation/quantitation with capillary gas chromatography equipped with selective detectors. The Netherlands’ Inspectorate for Health Protection has recently published in its 6th edition (1996) an extended collection of multiresidue methods for a multitude of pesticides: The Netherlands’ multiresidue method 1 (MRM-1) covers all pesticides that can be analyzed by capillary gas chromatography using selective detectors [electroncapture detector (ECD), nitrogen-phosphorus detector (NPD), flame photometric detector (FPD)] and increasingly full-scan ion trap mass spectrometric detection. Three submethods described in The Netherlands’ MRM-2 use high-performance liquid chromatography (HPLC) and postcolumn derivatization with fluorescence detection for N-methylcarbamate (including metabolites) and phenylurea pesticides or precolumn switching employing a pre-column packed with internal surface reversed-phase material for chlorophenoxy herbicides. The latter class of compounds is also analyzed more selectively by The Netherlands’ MRM-3 after derivatization with pentafluorobenzyl bromide (PFBBr) as esters. Derivatives of aromatic amines are covered by The Netherlands’ MRM-4 (two submethods) and by the German DFG S6 and S6-A methods, all of which use alkaline hydrolysis and steam distillation

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of the amines, followed by various derivatization procedures, and gas chromatography. The Luke method is a multiresidue method currently employed in the United States for enforcement of tolerances and import tolerances. Its cleanup includes mainly solid-phase extraction (SPE) cartridges of various selectivities with either gas or liquid chromatography of pesticides in separate fractions. However, because of the current emphasis on metabolites of toxicological concern, target methods are generally developed with an individual or closely related compounds in mind. Frequently, this involves simultaneous determination of the parent pesticide and its metabolites, which are generally of greater polarity (because they typically arise from oxidation or hydrolytic cleavage reactions).

51.3.2  Extraction Generally, extraction methodology must be developed such that nearly quantitative recovery of target analytes is obtained. At this stage of method development, the use of radiolabeled standards is invaluable because it allows for rapid determination of the percentage extractability by direct liquid scintillation counting (LSC). In later steps, the radiotracer is useful in determining efficiencies for each step of the proposed method. Traditional solvent extraction must take into account the chosen solvent’s water miscibility, ease of solvent concentration/removal, safety, and disposal costs. DFG S19 uses water/acetone (100 ml/200 ml) or acetonitrile/acetone for oily matrices and allows a sample size ranging from 10 g of very dry material (straw, herbs) to 100 g for watery matrices. The Netherlands’ multimethods use mainly acetone or ethyl acetate for extraction; other conventional extraction systems include methanol or acetonitrile. Some methods use combined extraction/hydrolysis steps either for deconjugation or to form common moiety products, which may be separated from the extraction mixture by steam distillation (e.g., The Netherlands’ MRM-4 and the DFG S6 for derivatives of aromatic amines). Several new extraction approaches are being developed such as supercritical fluid extraction (SFE) and pressurized liquid extraction (also known as accelerated solvent extraction, ASE). Typically, SFE gives very clean extracts with somewhat low recoveries and ASE gives good recoveries but the samples are exposed to high temperatures and pressures and require more extensive cleanup to remove co-extracted matrix components. The latter techniques allow only the extraction of relatively small amounts of samples (e.g., 5–10 g), which requires an increased emphasis on sample preprocessing and homogenization (to assure homogeneity) if market sample size amounts to 5–10 kg.

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SFE offers potential advantages for removing trace levels of target agrochemical analytes from various matrices. Of particular interest are the enhanced extraction rates obtained due to the high diffusivity of critical fluids. In addition to rapid extraction, improved penetration of the matrix with subsequent high recovery of bound residues is feasible (Hawthorne et al, 1992; Lira, 1988; Lopez-Avila and Dodhiwala, 1990). However, the success of commonly used fluid systems such as carbon dioxide or polar-modified carbon dioxide in binary mixtures is limited to, at best, moderately polar analytes. Because many factors can influence SFE efficiency (Erstfeld and Chen, 1998; Fahmy et al, 1993; Snyder et al, 1992, 1993), including pressure, temperature, fluid flow rate, extraction time, and modifier, considerable effort must be invested during method development before choosing SFE as the extraction technique. However, there is evidence that at least in some cases (as with chlorothalonil) it compares favorably to traditional approaches such as Soxhlet extraction (Erstfeld and Chen, 1998). Thorough descriptions of the instrumentation utilized by SFE are available in the literature (McNally et al, 1992; Riekkola et al., 1992). Excellent short reviews of SFE applications are those of Bond (1994) and King (1989). Supercritical fluid extraction has been reported to be successful in multiresidue analysis with pyrethroids (Argauer et al., 1997), organophosphates (Skopec et al., 1993), and other compound classes (Jones, 1996, 1997; King et al., 1993). In fact, even the hydrophobic avermectins are amenable to SFE techniques (Brooks and Uden, 1995). There is general agreement among researchers that although the scope of applicability for SFE may be limited, great advantages are provided in selected cases by the cleaner extracts obtained. Accelerated solvent extraction (ASE) is rapidly gaining acceptance as an alternative extraction approach with positive results in magnitude of residue, multiresidue, soil dissipation, and animal health studies (Stanek and Keller, 1998). ASE is based on the use of a variety of solvents under elevated temperatures and pressure. The higher temperatures involved accelerate the kinetics (as with Soxhlet) and the elevated pressure keeps the solvent in the liquid phase. The technique is amenable to automation (Ezzell, 1998). Traditionally, compounds that are difficult to extract from environmental matrices have been subjected to reflux conditions such as Soxhlet extraction. This approach is wasteful in time and solvent use. In fact, ASE compares favorably with Soxhlet and supercritical fluid extraction (David and Seiber, 1996; Frost et al., 1997; Lou et al., 1997). ASE is most effective with thermally stable low- or medium-polarity substrates. Extraction results are often better than those obtained with traditional methods (Conte et al., 1997; Ezzell et al., 1995). Sample preparation is quite

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simple (Richter et al., 1996), involving grinding and mixing of the soil followed by air drying or admixture with drying agents (e.g., sodium sulfate).

51.3.3  Cleanup of Extracts The primary extraction methodology can result in significant cleanup of the sample by separating the analyte from the bulk of interfering matrix components. However, primary extraction methods are designed to decrease sample bulk rather than to achieve complete purification. Therefore, additional steps are generally needed. The degree of purification ranges from none to very little with binary solvent systems (e.g., aqueous acetonitrile or methanol) to significant with solvent systems that allow homogeneous partition such as in the DFG S19 method, which uses a sequential system consisting of NaCl, water, acetone, ethyl acetate, and cyclohexane and results in rather clean extracts for watery crops, or with SFE or ASE, as discussed previously. Therefore, purification regimes must generally be instituted subsequent to the initial analyte separation. Traditional approaches to extract cleanup usually involve liquid/liquid partition. The most common technique focuses on differences in solubility (and polarity) of the matrix constituents. Thus, it is common to find methods in which (a) fatty components in matrices are removed by partition with nonpolar solvents (acetonitrile/hexane fat cleanup, extraction of aqueous extracts, or residues with dichloromethane) or (b) acidic or basic analytes are converted to water-soluble salts and the aqueous phase extracted with organic solvents to remove matrix components. Gel permeation chromatography (GPC) is an established cleanup technique that separates (size exclusion) high-molecular-weight compounds such as proteins, fats, and sugars from relatively low molecular weight compounds such as pesticides in animal and plant matrices. Typically, GPC involves injection of large-volume solutions of analyte (5 ml) via autosampler onto a low-pressure glass column containing polystyrene di-vinylbenzene beads conditioned with the appropriate solvent. Fractions of eluate are then collected and concentrated prior to further chromatographic analysis. A recent modification (Chambers, 1998) involves using a high-pressure steel column, lower injection volumes, and collection of smaller fractions, thus minimizing solvent use and concentration time. GPC is an integral part of multiresidue methods (Thier and Zeumer, 1987). A more efficient, but time-consuming method involves liquid/solid partition. For several decades, silica gel and florisil and size exclusion supports (gel permeation) have been used in liquid chromatography (LC). Thin-layer chromatography is far less effective. However, the advent of reverse-phase (RP) systems and high-efficiency SPE (and microextraction) cartridges has revolutionized the approaches to analyte purification. Whereas normal-phase

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(silica-based) supports could handle only organic solvents of medium to low polarity, reverse-phase systems can extract hydrophobic pesticide residues directly from aqueous solutions without involving significant amounts of organic solvents. The first report on applications of solid-phase extraction (SPE) with reverse-phase supports (Belardi and Pawliszyn, 1989) involved chemically modified fusedsilica fibers. There are now an increasing variety of other solid supports relying on ion exchange, size exclusion, and other physicochemical properties (Zhang et al., 1994). The technique has rapidly advanced, especially for the analysis of pesticide traces in systems such as surface and ground water (Balinova, 1993; Eisert and Levsen, 1995a; Field et al., 1997; Hatrik et al., 1994; Moore et al., 1995) and procymidone in wine (Urruty et al., 1997). SPE has also become an integral part of multiresidue methods (Analytical Methods for Pesticide Research, 1996; Barnabas et al, 1995; Nouri et al., 1995). A great advantage of SPE is that in many cases similar HPLC column supports are available (C8, C18, aminopropyl), which can be used to predict the chromatographic behavior of the analyte relative to potential interferences. Solid-phase extraction Empore disks are an alternative to SPE cartridges. These disks may eliminate certain cleanup procedures and further reduce the use of organic solvents. The disks may be extracted with small amounts of solvents directly in the autosampler vial (Field and Monahan, 1995, 1996), thus allowing for increased method automation. In general, SPE methodology is increasingly being incorporated into on-line systems (Marcé et al., 1995; Maris et al., 1985; Nielen et al., 1987) with an emphasis on polar pesticides such as diuron and bromacil in water samples (Parrilla et al., 1993; Sancho et al., 1997; Sennert et al., 1995).

51.3.4  Separation and Quantitation 51.3.4.1  Gas Chromatography and Mass Spectrometry For the past 40 years, gas chromatography (GC) has been the most widely utilized technique to analyze pesticide residues. Advances in chromatography, in particular, capillary column technology, have provided an increasing variety of thermally stable stationary phases, thus improving selectivity. The development of highly specific carbon, phosphorus, sulfur, and nitrogen detectors based on flame ionization and photometry (NPD, FPD) and on electron capture (ECD) culminated with the direct coupling of capillary GC columns with decreased carrier gas flow to mass spectrometers, in particular, those equipped with quadrupole analyzers (March, 1997). Thus, selectivity and sensitivity were improved tremendously. Large numbers of samples could thus be screened cheaply and efficiently once automation was introduced

Chapter | 51  Modern Approaches to Analysis of Pesticide Residues in Foods and the Environment

for sample injection. In typical applications, such a GC–MS system can provide quantitation of 20–40 samples overnight, including full- or selected-ion spectra on several components. In fact, GC–MS is already being used extensively for the U.S. Department of Agriculture Pesticide Data program, as exemplified by the routine simultaneous determination of diphenyl amine, o-phenylphenol, and propargite in apples (Yu et al., 1997). Other significant developments in GC-MS included the following: 1. The use of high-resolution mass spectrometry (HRMS) with magnetic sector instruments, which led to limits of detection in the femtogram range, especially for analytes that contain heteroatoms with significant mass defects (e.g., chlorine) and that are prone to give simple spectra on negative chemical ionization (NCI). 2. The introduction of the ion trap mass spectrometer (ITD), which resulted in better sensitivities in the fullscan mode, thus providing improved identification power in combination with nontarget multiresidue enforcement methods. 3. The introduction of multiple MS capabilities either by means of a row of quadrupoles (triple-quads), providing MS/MS in space; or by the use of ion trap mass spectrometers, allowing MSn (multiple MS experiments) in time. The MS/MS or MSn techniques result quite often in better selectivity and thus in improved sensitivity compared to the single-quadrupole or early ion trap techniques, whereas the price for the instruments is still much lower than that for the high-resolution MS instruments. Thus, the versatility of these instruments, especially of the ion trap mass spectrometers equipped either with an external ionization source (which allows negative and positive chemical ionization) or with internal ionization (which seems to give better ion yields on electron impact and positive chemical ionization), has resulted in a tremendous improvement in the limit of detection for many analytes that can be chromatographed by capillary GC. Multiple MS thus serves as an additional “cleanup” technique, allowing for very low detection limits of analytes in matrix. For example, nanogram per liter detection of several pesticides in aquatic matrices has been reported using tandem GC–MS (Boyd-Boland et al., 1996; Rossi et al., 1997; Steen et al., 1997). That GC-MS/MS is rapidly becoming the preferred technique for pesticide analysis (Feigel, 1997) is due in part to the added capability for confirmation of identity, which allows for elimination of the false positives often detected in GC–MS analysis (de Cruz et al., 1996; Schachterle and Feigel, 1996). Furthermore, the possibility of conducting two or more MSn experiments in the search for ions not present in co-eluting interferences greatly simplifies the extraction and purification process. GC–MS has proven useful in multiresidue methods,

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sometimes dealing with mixtures containing in excess of 100 compounds, in a very cost-effective manner (Fillion et al., 1995; Liao et al., 1991). In accordance with present trends toward automation of pesticide analysis, it may be expected that in-line techniques coupling solid-phase microextraction of pesticide mixtures with GC–MS (Eisert and Levsen, 1995b) will be increasingly reported.

51.3.4.2  Liquid Chromatography and Mass Spectrometry A major shortcoming of GC-MS techniques is their inability to analyze samples of low volatility, high polarity, or thermal instability. Because regulatory requirements focused only on the parent pesticide, GC–MS could accomplish the required goals for the great majority of commercial products. However, in the past decade the need for quantitative data on degradation products and metabolites has increased and now the complete expression for “toxic residues” routinely includes one or more metabolites in addition to the pesticide itself. A multitude of derivatization techniques (Lunn and Hallwig, 1998) is available to convert hydroxy, carboxy, amino, and other polar derivatives to entities amenable to GC–MS. The derivatization approach is expensive and often unreliable in view of the variety of matrices and co-extractives involved. Furthermore, the analytes of interest are often obtained in aqueous fractions incompatible with the majority of available derivatizing agents. Liquid chromatography (LC) can successfully handle polar, nonvolatile compounds but, as a residue technique, it has developed more slowly than GC, and, mainly due to the lack of element-specific detectors, LC has not reached the wide applicability of gas chromatographic methods. Although detectors utilizing ultraviolet absorbance, fluorescence (with/without postcolumn derivatization), electrochemical properties, and refractive index have proven useful, the advent of combined liquid chromatography and mass spectrometry (LC–MS) has provided renewed and strong interest in LC as a residue analytical tool. The union of the two analytical techniques in LC-MS combines an instrument that operates in the condensed phase with one that operates at reduced pressures. Thus, for LC, it is the mobile phase and flow rate that affects separation of analytes; for MS, it is the ionization mode and factors affecting the production and transport of gas-phase ions into the analyzer that are important. In the 1980s, several approaches evolved toward finding a practical application for LC–MS: First were the moving belt interface and direct liquid introduction, which were later replaced by particle beam and thermospray technologies (Cairns and Siegmund, 1990). These methods have been applied to a great variety of analytical problems involving pesticides, their metabolites, and

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conjugates (Brown, 1990). In particular, it is worth noting that LC–MS proved itself as a viable technique with the successful analysis of thermally labile sulfonylurea herbicides (Shalaby and George, 1990). These compounds are used at very low application rates so the use of thermospray techniques established the increased sensitivity of LC–MS as compared to other methods. In spite of interface improvements such as thermospray, complications in the use of LC–MS were common and based on the introduction of a fluid stream into a vacuum system. At present, the best results are being obtained with ion sources developed over the past decade that operate at or near atmospheric pressure. Atmospheric pressure chemical ionization (APCI) relies on nebulization of the solvent stream followed by thermal evaporation. Thus, the mixture is ionized in the vapor phase and the reactant ions are formed from the components present in the LC eluent. Chemical ionization utilizes solvent molecules as the “reagent gas.” APCI can handle high flow rates and high electrolyte concentrations. The second widely utilized atmospheric pressure technique is electrospray ionization (ESI). Here the sample has to be in an ionized form in solution. Neutral samples can be converted to ions by adjustment of pH or by addition of electrolytes (such as ammonium acetate) to form ion-molecule complexes. The influence of the species utilized for ionpair processes in LC is quite important as evidenced in the analysis of diquat and paraquat (Startin et al., 1998). The sample solution is then dispersed into an electrically charged aerosol. The target ions are separated from the droplet interface by the action of an electric field at the surface of the charged droplet, which, upon partial solvent removal, develops a substantially smaller cross section. The gas-phase ions are then transported to the mass analyzer. Because electrospray ionization can be accomplished at low temperatures, this technique is ideal for thermally labile compounds. ESI generally works best at low flow rates and with lower concentrations of electrolytes. Both APCI and ESI are quite sensitive to eluent composition and to matrix effects. Thus, extensive method development is required in the early stages of analysis to optimize sensitivity, selectivity, and compatibility with matrix components and eluent solvents. In spite of these drawbacks, APCI and ESI are rapidly becoming the LC–MS techniques of choice. As with GC–MS, tandem (MS/MS), or MSn, methods for LC have proliferated (Gilbert et al., 1995). The great advantage of the MS/MS technique using triple-quadrupole instruments (MS/MS-in-space) and of the MS– (-in-time) technique using ion trap technology lies in the gain of selectivity using the “multiple-reaction mode” (MRM). In the case of APCI, protonated quasi-molecular parent ions are selected either by passing the first quadrupole or by eliminating all other ions in the ion trap, and then fragmented further by collision-induced dissociation in the middle quadrupole

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or in the trap by applying energy. The daughter ion fragments formed are then filtered by the third quadrupole or by the ion trap and yield highly selective MS signals. The majority of pesticides and metabolites can be detected by LC–MS/MS in concentrations greater than 10 ng/ml with high selectivity, allowing much abbreviated cleanup procedures. However, although the current view is that MS/MS techniques can largely eliminate cleanup steps, the issues and strategies that have historically been important in sample preparation and analysis remain fundamentally important to reliable LC–MS/MS. The “dilute and shoot” approach is valid in some cases where analyte concentrations are high and matrix components do not coelute or otherwise interfere with ionization. Contamination of HPLC column and MS source by matrix components, however, may cause variation and drift of the LC-MS signals and thus hamper unattended automated analysis of crude extracts. This can be avoided to a certain extent by introducing precolumn-switching techniques or automated on-line SPE cleanup steps. An obvious approach to bypass matrix effects is to prepare and analyze the quantitation standards in solutions containing matrix. However, the EC and U.S. guidelines diverge on this point. Whereas in Europe the use of standard in-matrix calibrants is encouraged, the EPA does not generally allow it. As with all mass spectrometric techniques, the use of isotopically labeled internal standards results in an improved reliability and ruggedness of methods. This approach is very well established in pharmaceutical analysis, but nonradioactivelabeled tracer compounds are available for only a limited number of pesticides and metabolites. Use of internal 13 C- or 2H-labeled standards is of particular usefulness when extraction techniques result in partial losses of analytes, as in the case with chlorinated anilines (Hurlbut et al., 1998), which react with matrix components, and of pentachlorophenol (Gremaud and Turesky, 1997). Analytical methods using GC or LC-MS/MS and MSn techniques are per se target methods (i.e., tailored to detect with high selectivity and sensitivity one or few analytes). Thus, the use of MS/MS has limitations for enforcement methods that should ideally cover a multitude of pesticides and relevant metabolites. This is especially true for LC–MS due to the limited separation power of HPLC in comparison to the much higher peak capacity obtained by capillary GC. This disadvantage can be overcome by two approaches, as follows. First, if several rugged automated short LC–MSn methods analyze sample extracts after a general cleanup procedure consecutively for several groups or classes of compounds, this would allow unattended screening of one extract for many analytes. As sample extraction and cleanup are time-consuming steps, this approach allows increased sample throughput for enforcement purposes. Second, a technique called data-dependent full-scan MS/MS can be implemented with ion trap mass spectrometers.

Chapter | 51  Modern Approaches to Analysis of Pesticide Residues in Foods and the Environment

This application first screens in the full-scan mode the mass spectral information for compounds eluting from the HPLC column. As soon as ions formed from a potential analyte are detected in a defined retention time window, the ion trap switches to a predetermined MS/MS method for improved selectivity, thus providing additional information and confirmation. A combination of these two approaches could very well result in the future in reliable and affordable multiresidue methods for pesticides and relevant metabolites not covered by GC-based multimethods.

51.3.5  Immunoassay Techniques The use of immunoassays (IAs) as pesticide residue analytical methods represents a radical departure from the more conventional chromatographic approaches. Immunoassays utilize antibodies that have been prepared in animals (commonly rabbits, mice, or sheep) to a particular pesticide or family of pesticides. These molecules are too small to elicit immune responses by themselves, but, upon coupling of a chemical analog of the pesticide to a carrier (usually a protein), the “conjugate” may evoke production of antibodies. For the method to be successful, these antibodies must be able to bind selectively to the free pesticide. The key steps in development of an antibody test are as follows (Gee et al., 1995; Harris et al., 1998): 1. Synthesis of a pesticide (or derivative), coupled to a suitable carrier protein for immunization. 2. Immunization of rabbits, mice, and/or other species; preparation and purification of antibodies. 3. Development of initial immunoassay using pesticide standards; checking assay sensitivity and specificity. 4. Assessment of assay performance with water and soil matrices in laboratory-spiked and field samples. 5. Formatting of methods as prototype kits, stabilization and stability trials on components and prototypes. 6. Field trials of kits and training workshops. The advantages associated with IA include low detection limits and high analyte selectivity. Because sample preparation is minimal, the methods allow for high throughput, thus increasing cost effectiveness. These advantages have been extensively reviewed in the literature (Harris et al., 1995, 1998). A key limitation in the development of IA methodology is the longer time required when compared to traditional instrumental methods. Specifically, the selection of the target analyte analog (hapten) is critical for the production of the high-affinity antibodies required for high selectivity and sensitivity. The functional group used for protein coupling should not mask the key structural feature(s) of the target analytes so that the immune system of the host animal can recognize it. Hapten design and synthesis have been extensively reviewed (Goodrow et al., 1990, 1995).

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The coupling (conjugation) of the hapten with a carrier protein must result in a product that is of adequate solubility and stability under the reaction conditions and that contains the appropriate functional groups (Brinckley, 1992; Erlanger, 1980). The unreacted products are then separated from the conjugate by dialysis, gel filtration, or other methods. Animal immunization to obtain polyclonal or monoclonal antibodies is conducted under conditions that enhance the immune response (Harlow and Lane, 1988; Tijssen, 1985). The affinity of the antibodies for the immunizing hapten is then evaluated to determine antibody titers against it. Inhibition experiments are then conducted to determine the potential for each target analyte toward inhibition of binding between the hapten and the antibody. Validation of the IA method is conducted after matrix effects and the influence of pH, salts, solvents, and other components are identified. As with other animal methods, validation involves determination of fortified recoveries in the appropriate matrices. Schneider et al., (1995) detail troubleshooting procedures that may be considered during method development and validation. Numerous examples of successful IA methods are reported for organophosphates, pyrethroids, triazines, urea herbicides, and other compound classes (Gee et al., 1995; Harris et al., 1998). For the analysis of specific agrochemicals, commercially available immunoassay kits are generally more cost effective than traditional instrumental analysis. This factor makes the continuing development of IA methods attractive, in particular, for use in monitoring studies and for pesticide analysis in developing countries. However, substantial purification of analyte from matrix is required before IA methods can be applied.

51.3.6  Capillary Electrophoresis Electrophoresis refers to the migration of electrically charged species when dissolved in an electrolyte through which an electric current is passed. Capillary electrophoresis (CE) combines a variety of modern analytical techniques with a wide range of applications, including the analysis of biopolymers [such as deoxyribonucleic acid (DNA), proteins, peptides], natural products, pharmaceuticals and drugs, and fine chemicals, including agrochemicals. CE makes use of various separation modes with distinct ranges of application, such as the following: 1. Free-solution capillary electrophoresis (FSCE), which is mainly used for the separation of ions based on differences in their charge-to-mass ratios, whereby the pH controls the dissociation and protonation of functional groups on the analytes. 2. Micellar electrokinetic capillary chromatography (MECC or MEKC), which uses relatively high levels of ionic surfactants forming micelles. The separation of neutral analytes is based on the hydrophobic interaction with the

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micelles that migrate in the capillary. The use of chiral cyclodextrins provides a relatively cost effective and powerful method for enantioselective separations. 3. Capillary electrochromatography (CEC) is a fusion of liquid chromatography and capillary electrophoresis, where the capillary is packed with stationary phase similar to those used in liquid chromatography, and the flow of the mobile phase is caused by the electroosmotic flow (EOF) between the electrodes. The selectivity of the separation depends on partition between the stationary and mobile phases. Capillary electrophoresis has to be considered for the analysis of polar and charged analytes and thus follows the trend in pesticide chemistry to use more hydrophilic watersoluble active substances and for the requirement to include polar metabolites into the residue definition. For routine residue analysis, however, CE is generally only considered when conventional chromatographic approaches such as capillary GC, HPLC, and LC-MS fail to provide straightforward solutions, for the following reasons. First, expertise and instrumentation in GC, HPLC, and LC-MS are more readily available in residue research, contract, and enforcement laboratories, and the pressure to use CE for residue analysis only exists when the other techniques do not provide rational and cost-effective methods. Second, there are several aspects of CE that do not facilitate its use in residue analysis: 1. Whereas CE provides excellent sensitivity for many applications, the limitation in sample size, which is in the nanoliter range, results in insufficient overall sensitivity. Injection size can be increased by the use of a process called “stacking” where the analytes are concentrated in a sample zone prior to the chromatographic separation, thus reducing the starting peak width. Miniaturization of extraction and cleanup techniques may provide a means to obtain decreased final extract volumes. However, in residue analysis, there is a limit to decreasing the original sample size. 2. The most frequently used detection methods in CE are ultraviolet (UV) absorbance or UV diode array. Very short path lengths, however, again limit the sensitivity obtained by CE. On the other hand, the use of a mass spectrometric detector (quadrupole, triple-quadrupole, ion trap, or time-of-flight mass spectrometer, MS/MS techniques), predominantly with ESI sources, provides good sensitivity and high selectivity. The presence of electrolytes and surfactants in the case of MECC, however, limits applications and overall performance. In summary, with the current trend away from highly lipophilic pesticides of great environmental persistence toward more polar active substances and the necessity to detect polar metabolites, there is an increasing number of potential applications in pesticide residue analysis for CE.

51.4  Summary Regulatory requirements that address pesticide concentrations in foods and environmental matrices have significantly accelerated the development of faster, less costly, and more sensitive and specific analytical techniques. Extraction techniques utilizing modern approaches such as supercritical fluid and accelerated solvent extractions allow for use of smaller sample sizes and solvent volumes. Cleanup of extracts can now be accomplished with commercial chromatography products and the process can be automated to address large sample numbers. The most dramatic improvements have taken place in the analytical instrumentation and automation options available. In particular, the development of multiple ionization techniques and secondary ion production in mass spectrometry have advanced detection limits and improved selectivity. Developments in immunoassay-based analysis promise to provide low detection limits and high analyte selectivity coupled with relatively low cost.

References Argauer, A. J., Eller, K. I., Pfeil, R. M., and Brown, R. J. (1997). Determining ten synthetic pyrethroids in lettuce and ground meat by using ion-trap mass spectrometry and electron capture gas chromatography. J. Agric. Food Chem. 45, 180–184. Balinova, A. (1993). Solid-phase extraction followed by high-performance liquid chromatographic analysis for monitoring herbicides in drinking water. J. Chromatogr. 643, 203–207. Barnabas, I. J., Dean, J. R., Fowlis, I. A., and Owen, S. P. (1995). Automated determination of s-triazine herbicides using solid-phase microextraction. J. Chromatogr. 705, 305–312. Belardi, R. P. and Pawliszyn, J. B. (1989). The application of chemically modified fused silica fibers in the extraction of organics from water matrix samples and their rapid transfer to capillary columns. Water Pollut. Res. J. Can. 24, 179–189. Bond, P. M. (1994). Supercritical fluid chromatography—Finding a place in the chromatographer’s arsenal? Pestic. Sci. 41, 369–374. Boyd-Boland, A. A., Magdic, S., and Pawliszyn, J. (1996). Simultaneous determination of 60 pesticides in water using solid phase microextraction and gas chromatography-mass spectrometry. Analyst 121, 929–938. Brinkley, M. (1992). A brief survey of methods for preparing protein conjugates with dyes, haptens and cross-linking reagents. Bioconj. Chem. 3, 2. Brooks, M. W., and Uden, P. C. (1995). Determination of abamectin from soil and animal tissue by SFE and fluorescence detection. Pestic. Sci. 43, 141–146. Brown, M. A. (1990). (ed.), “Liquid Chromatography/Mass Spectrometry,” ACS Symposium Series 420. Am. Chem. Soc., Washington, DC. Cairns, T., and Siegmund, E. G. (1990). The development of liquid chromatography/mass spectrometry. In “Liquid Chromatography/Mass Spectrometry” (M. A. Brown, ed.), pp. 1–15, ACS Symposium Series 420. Am. Chem. Soc., Washington, DC. Chambers, J. (1998). Miniaturized GPC for the effective cleanup of crop extracts containing pesticide residues. In “Ninth International Congress of Pesticide Chemistry (IUPAC),” Vol. 2, 7A-046.

Chapter | 51  Modern Approaches to Analysis of Pesticide Residues in Foods and the Environment

Conte, E., Milani, R., Morali, G., and Aballe, F. (1997). J. Chromatogr. A 765, 121–125. David, M. D. and Seiber, J. N. (1996). Comparison of extraction techniques, including supercritical fluid, high-pressure solvent, and Soxhlet, for organophosphorus hydraulic fluids from soil. Anal. Chem. 68, 3038–3044. de Cruz, I., Lacroix, G., Mougin, C., and Grolleau, G. (1996). Residues of chlorinated pesticides in eggs of the gray heron: Contribution of capillary gas chromatography ion-trap spectrometry. J. High Resolut. Chromatogr. 19, 62–64. EC Directorate General for Agriculture (1998). “Guidance Document on Residue Analytical Methods.” 8064/VI/97-rev 4 15/12/98. Eisert, R. and Levsen, K. (1995a). Determination of organophosphorus, triazine and 2,6-dinitroaniline pesticides in aqueous samples via solidphase microextraction (SPME) and gas chromatography with nitrogenphosphorus detection. Fresenius’ J. Anal. Chem. 351, 555–562. Eisert, R. and Levsen, K. (1995b). Determination of pesticides in aqueous samples by solid-phase microextraction in-line coupled to gas chromatography-mass spectrometry. J. Am. Soc. Mass Spectrom. 6, 1119–1130. Erlanger, B. F. (1980). The preparation of antigenic hapten-carrier conjugates: A survey. Methods Enzymol. 70, 85. Erstfeld, K. M and Chen, C. Y. (1998). Comparison of supercritical fluid and Soxhlet extraction of chlorothalonil from cranberry bog soils. J. Agric. Food Chem. 46, 499–503. Ezzell, J. (1998). The use of SW-846 Method 3545 for automated extraction of environmental samples. Am. Environ. Lab., 24–25. Ezzell, J., Richter, B., Felix, D., Black, S., and Meikle, J. (1995). A comparison of accelerated solvent extraction with conventional solvent extraction for organophosphorus pesticides and herbicides. LC-GC, 390–398. Fahmy, T. M., Paulaitis, M. E., Johnson, D. M., and McNally, M. E. P. (1993). Modifier effects in the supercritical fluid extraction of solutes from clay, soil, and plant materials. Anal. Chem. 65, 1462–1469. Feigel, C. (1997). GC/MS/MS: Method of choice for pesticide residue analysis. Inside Laboratory Management, 34–35. Field, J. A. and Monahan, K. (1995). In-vial derivatization and Empore disk elution for the quantitative determination of the carboxylic acid metabolites of dacthal in groundwater. Anal. Chem. 34, 3357–3362. Field, J. A. and Monahan, K. (1996). Chlorinated acid herbicides analysis in water by strong anion-exchange disk extraction and in-vial elution and derivatization. J. Chromatogr., A 741, 85–90. Field, J. A., Reed, R. L., Sawyer, T. E., and Martinez, M. (1997). Diuron and its metabolites in surface and ground water by solid phase extraction and in-vial elution. J. Agric. Food Chem 45, 3897–3902. Fillion, J., Hindle, R., Lacroix, M., and Selwyn, J. (1995). Multiresidue determination of pesticides in fruits and vegetables by gas chromatographymass-selective detection and liquid chromatography with fluorescence detection. J.—Assoc. Off. Anal. Chem. 78, 1252–1266. Frost, S., Dean, J., Evans, K., Harradine, K., Cary, C., and Comber, M. (1997). Extraction of hexaconazole from weathered soils: A comparison between Soxhlet extraction, microwave assisted extraction, supercritical fluid extraction and accelerated solvent extraction. Anal. Chem. 69, 2171–2180. Gee, S. J., Hammock, B. D., and Skerritt, J. H. (1995). Diagnostics for plant agrochemicals—A meeting of chemistry and immunoassay. In “New Diagnostics in Crop Sciences” (J. H. Skerritt and R. Appels, eds.), pp. 243–276. CAB, International, Oxford. Gilbert, J. D., Olah, T. V., and McLaughlin, D. A. (1995). Biochemical and biotechnological applications of electrospray ionization mass spectrometry. In “ACS Symposium Series 619” (A. P. Snyder, ed.), pp. 330–350. Am. Chem. Soc, Washington, DC.

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Goodrow, M. H., Harrison, R. O., and Hammock, B. D. (1990). Hapten synthesis, antibody development, and competitive inhibition enzyme immunoassay for s-triazine herbicides. J. Agric. Food Chem. 38, 990. Goodrow, M. H., Sanborn, J. R., Stoutamire, D. W., Gee, S. J., and Hammock, B. D. (1995). Strategies for immunoassay hapten design. In “Immunoanalysis of Agrochemicals: Emerging Technologies” (J. O. Nelson, A. E. Kara, and R. B. Wong, eds.), pp. 119–126, ACS Symposium Series 586. Am. Chem. Soc, Washington, DC. Gremaud, E. and Turesky, R. J. (1997). Rapid analytical methods to measure pentachlorophenol in wood.. J. Agric. Food Chem. 95, 1229–1233. Harlow, E. and Lane, D. (1988). “Antibodies: A Laboratory Manual.” Cold Spring Harbor Laboratory Press, Cold Spring Harbor, NY. Harris, A. S., Lucas, A. D., Kramer, R M., Marco, M. R, Gee, S. J., and Hammock, B. D. (1995). Use of immunoassay for the detection of urinary biomarkers of exposure. In “New Frontiers in Agrochemical Immunoassay” (D. A. Kurtz, J. H. Skerritt, and L. Stanker, eds.), p. 217. Assoc. Official Anal. Chem., Arlington, VA. Harris, A. S., Wengatz, I., Wortberg, M., Kreissig, S. B., Gee, S. J., and Hammock, B. D. (1998). Development and application of immunoassays for biological and environmental monitoring. In “Multiple Stresses in Ecosystems” (J. J. Cech, B. W. Wilson, and D. G. Crosby, eds.), pp. 135–153. Lewis Publishers, CRC Press, NY. Hatrik, S., Lehotay, J., and Tekel, J. (1994). Simultaneous HPLC determination of phenylurea herbicides and their corresponding anilines in water after online preconcentration. J. High Resolut. Chromatogr. 17, 756–758. Hawthorne, S. B., Miller, D. J., and Langefeld, J. J. (1992). Advances in analytical supercritical fluid extraction (SFE). In “Hyphenated Techniques in Supercritical Fluid Chromatography and Extraction” (K. Jinno ed.) Chap. 12. Elsevier, Amsterdam. Hurlbut, D. B., Johnston, J. J., Daniel, S. R., and Tawara, J. (1998). Gas chromatography/mass spectrometry method for the quantitation of 3-chlorop-toluidine hydrochloride in birds using a deuterated surrogate. J. Agric. Food Chem. 46, 4610–4615. Inspectorate for Health Protection (1996). “Analytical Methods for Pesticide Research,” 6th ed. Inspectorate for Health Protection, Rijswijk, The Netherlands. Jenke, D. R. (1996a). Chromatographic method validation: A review of current practices and procedures. I. General concepts and guidelines. J. Liq. Chromatogr. Relat. Technol. 19, 719–736. Jenke, D. R. (1996b). Chromatographic method validation: A review of current practices and procedures. II. Guidelines for primary validation parameters. J. Liq. Chromatogr. Relat. Technol. 19, 737–757. Jenke, D. R. (1996c). Chromatographic method validation: A review of current practices and procedures. III. Ruggedness, re-validation, and system suitability. J. Liq. Chromatogr. Relat. Technol. 19, 1873–1891. Jones, F. W. (1996). Multi-residue analysis of pesticides in wool wax and lanolin using gel permeation and gas chromatography. J. Agric. Food Chem. 44, 3197–3201. Jones, F. W. (1997). SFE as a cleanup technique for gas chromatographic analysis of pesticides in wool wax. J. Agric. Food Chem. 45, 2569–2572. King, J. W. (1989). Fundamentals and applications of supercritical fluid extraction in chromatographic science. J. Chromatogr. Sci. 27, 355–364. King, J. W., Hopper, M. L., Lutchefeld, R. G., Taylor, S. L., and Orton, W. L. (1993). Optimization of experimental conditions for the supercritical carbon dioxide extraction of pesticides residues from grains. J.—Assoc. Off. Anal. Chem. Int. 76, 857–864.

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Liao, W., Joe, T., and Cusick, W. G. (1991). Multi-residue screening method for fresh fruits and vegetables with gas chromatographic/mass spectrometric detection. J.—Assoc. Off. Anal. Chem. 78, 554–565. Lira, C. T. (1988). Physical chemistry of supercritical fluids. In “Supercritical Fluid Extraction and Chromatography” (B. A. Charpentier and R. S. Michael, eds.), Chap. 1. Am. Chem. Soc, Washington, DC. Lopez-Avila, V. and Dodhiwala, N. S. (1990). Supercritical fluid extraction and its application to environmental analysis. J. Chromatogr. Sci. 28, 468–476. Lou, X., Janssen, H., and Cramers, C. (1997). Parameters affecting the accelerated solvent extraction of polymeric samples. Anal. Chem. 69, 1598–1603. Lunn, G. and Hallwig, L. C. (1998). “Handbook of Derivatization Reactions for HPLC.” Wiley, New York. Marcé, R. M., Prosen, H., Crespo, C., Calull, M., Borrull, F., and Brinkman, U. A. Th. (1995). On-line trace enrichment of polar pesticides in environmental waters by reversed-phase liquid chromatography–diode array detection–particle beam mass spectrometry. J. Chromatogr. A 696, 63–74. March, R. E. (1997). An introduction to quadrupole ion trap mass spectrometry. J. Mass Spectrom 32, 351–369. Maris, F. A., Geerdink, R. B., Frei, R. W, and Brinkman, U. A. Th. (1985). On-line trace enrichment for improved sensitivity in liquid chromatography with direct liquid introduction mass spectrometric detection. J. Liq. Chromatogr. 323, 113–120. McNally, M. E. P., Deardorff, C. M., and Fahmy, T. M. (1992). Supercritical fluid extraction—new directions and understandings. In “Supercritical Fluid Technology” (F. Bright and M. E. McNally, eds.), Chap. 12. Am. Chem. Soc, Washington, DC. Moore, K. M., Jones, S. R., and James, C. (1995). Multi-residue analytical method for diuron and carbamate pesticides in water using solidphase extraction and liquid chromatography-mass spectrometry. Water Res 29, 1225–1230. Nielen, M. W. F., Valk, A. J., Frei, R. W., Brinkman, U. A. Th., Mussche, Ph., De Nijs, R., Ooms, B., and Smink, W. (1987). Fully automated sample handling system for liquid chromatography based on pre-column technology and automated cartridge exchange. J. Chromatogr. 393, 69–83. Nouri, B., Fouillet, B., Toussiant, G., Chambon, P., and Chambon, R. (1995). High-performance liquid chromatography with diode array detection for the determination of pesticides in water using automated solid-phase extractions. Analyst 120, 1133–1136. Parrilla, P., Martinez-Vidal, J. L., and Fernandez-Alba, A. R. (1993). Optimization of the separation, isolation and recovery of selected pesticides in water samples by solid-phase extraction and HPLC photodiode array detection. J. Liq. Chromatogr. 16, 4019–4029. Ragsdale, N. N. (1998). Impact of US Food Quality Protection Act on pesticide research priorities. In “Ninth International Congress of Pesticide Chemistry (IUPAC),” Vol. 2, 8A-006. Richter, B., Jones, B., Ezzell, J., and Porter, N. (1996). Accelerated solvent extraction: A technique for sample preparation. Anal. Chem. 68, 1033–1039. Riekkola, M. L, Manninen, P., and Hartonen, K. (1992). SFE, SFE/GC and SFE/SFC: Instrumentation and applications. In “Hyphenated Techniques in Supercritical Fluid Chromatography and Extraction” (K. Jinno, ed.), Chap. 14. Elsevier, Amsterdam. Rossi, D., Hoffman, K., Janiczek-Dolphin, N., Bockbrader, H., and Parker, T. (1997). Tandem-in-time mass spectrometry as a quantitative bioanalytical tool. Anal. Chem. 69, 4519–4523. Sancho, J. V., Hidalgo, C., and Hernández, F. (1997). Direct determination of bromacil and diuron residues in environmental water samples

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by coupled-column liquid chromatography and large-volume injection. J. Chromatogr., A 761, 322–326. Schachterle, S. and Feigel, C. (1996). Pesticide residue analysis in fresh produce by gas chromatography-tandem mass spectrometry. J. Chromatogr., A 754, 411–422. Schneider, P., Gee, S. J., Kreissig, S. B., Harris, A. S., Kramer, P., Marco, M. P., Lucas, A. D., and Hammock, B. D. (1995). Troubleshooting during the development and use of immunoassays for environmental monitoring. In “New Frontiers in Agrochemical Immunoassay” (D. A. Kurtz, J. H. Skerritt, and L. Stanker, eds.), pp. 103–130. Assoc. Official Anal. Chem., Arlington, VA. Sennert, S., Volmer, D., Levsen, K., and Wiinsch, G. (1995). Multiresidue analysis of polar pesticides in surface and drinking water by on-line enrichment and thermospray LC-MS. Fresenius’ J. Anal. Chem. 351, 642–649. Shalaby, L. M., and George, S. W. (1990). Multiresidue analysis of thermally labile sulfonylurea herbicides in crops by liquid chromatography/ mass spectrometry. In “Liquid Chromatography/Mass Spectrometry” (M. A. Brown, ed.), pp. 75–92, ACS Symposium Series 420. Am. Chem. Soc, Washington, DC. Skopec, Z. V., Clark, R., Harvey, P. M. A., and Wells, R. J. (1993). Analysis of organophosphorus pesticides in rice by supercritical fluid extraction and quantitation using an atomic emission detector. J. Chromatogr. Sci. 31, 445–449. Snyder, J. L., Grob, R. L., McNally, M. E., and Oostdyk, T. S. (1992). Comparison of supercritical fluid extraction with classical sonication and Soxhlet extractions of selected pesticides. Anal. Chem. 64, 1940–1946. Snyder, J. L., Grob, R. L., McNally, M. E., and Oostdyk, T. S. (1993). The effect of instrumental parameters and soil matrix on the recovery of organochlo-rine and organophosphate pesticides from soil using supercritical fluid extraction. J. Chromatogr. Sci 31, 183–191. Specht, W., Pelz, S., and Gilsbach, W. (1995). Gas-chromatographic determination of pesticide residues after cleanup by gel-permeation chromatography and mini-silica gel chromatography. Fresenius’ J. Anal. Chem. 353, 183–190. Stanek, M., and Keller, G. (1998). Determination of pesticide residues using accelerated solvent extraction. In “Ninth International Congress of Pesticide Chemistry (IUPAC),” Vol. 2, 7A-034. Startin, J. R., Hird, S. J., Jones, A., and Hill, A. R. C. (1998). Analysis of residues of paraquat and diquat in plant and animal tissues by LC-MS. In “Ninh International Congress of Pesticide Chemistry Book of Abstracts,” Vol. 2, 7A-007. Steen, R. J. C. A., Freriks, I. L., Cofmo, W. P., and Brinkman, U. A. Th. (1997). Large-volume injection gas chromatography–ion trap tandem mass spectrometry for the determination of pesticides in the marine environment at low ng/1 level. Anal. Chem. Actal 353, 153–163. Thier, H. P. and Zeumer, H. (1987). “Manual of Pesticide Residue Analysis,” DFG Pestic. Comm, Weinheim. Tijssen, P. (1985). “Laboratory Techniques in Biochemistry and Molecular Biology, Practice and Theory of Enzyme Immunoassays.” Elsevier, Amsterdam. Urruty, L., Montury, M., Braci, M., Founder, J., and Dournel, J. M. (1997). Comparison of two recent solventless methods for determination of procymidone residues in wine. J. Agric. Food Chem. 45, 1519–1522. U.S. Environmental Protection Agency (EPA) (1996). “Residue Chemistry Test Guidelines,” OPPTS 860 Series. Yu, L., Schoen, R., Dunkin, A., Firman, M., and Cushman, H. (1997). Rapid identification and quantitation of diphenylamino, o-phenylphenol and propargite residues on apples by GC–MS. J. Agric. Food Chem. 45, 748–752. Zhang, Z., Yang, M. J., and Pawliszyn, J. (1994). Solid-phase microextraction. Anal. Chem. 66, 844–853.

Chapter 52

Worker Exposure: Methods and Techniques Graham Chester OCCUBEX RA Limited, Hampshire, UK

52.1  Introduction Pesticides are biologically active compounds, which may pose a health risk to agricultural workers during or after their use. Operators involved in handling, dispensing, and applying pesticides and postapplication crop re-entry workers will be exposed to these compounds through different routes and to varying extents. It is essential both for stewardship and regulatory approval purposes that the possible health risk associated with this exposure is assessed using quantitative information on the toxicological hazard and the amount of exposure. This chapter presents methods by which exposure can be determined. It is not intended to be a complete review of the literature or to provide detailed guidance on how to measure operator exposure or conduct a field exposure study but rather a summary of current “state-of-the-art” principles and methodology involved in measuring exposure of agricultural workers to pesticides. The processes by which such exposure data are used and interpreted are dealt with elsewhere in this handbook.

52.2  Routes of exposure Agricultural workers involved in the use of pesticides and postapplication crop re-entry activities may be exposed to pesticides via the skin, by inhalation, or by accidental oral ingestion. Exposures via the first two routes are usually determined separately, but little attention is paid to oral ingestion because it is difficult to estimate or measure. Exposure is usually greatest by the dermal route, although inhalation can be an important route for pesticides that have significant vapor pressures, are applied in confined Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

spaces, or have an application technique that generates a significant proportion of respirable or inhalable particles.

52.2.1  Dermal Exposure In practice, two measurements or estimations are usually made for all work activities associated with the use of pesticides: 1. Potential dermal exposure – the total amount of pesticide coming into contact with the protective clothing, work clothing, and skin. 2. Actual dermal exposure – the amount of pesticide coming into contact with the bare (uncovered) skin and the fraction transferring through protective and work clothing or via seams to the underlying skin, which is therefore available for percutaneous absorption. The biological availability or absorption of a pesticide via the dermal route of exposure is a property of the formulated product and the diluted material and is a separate subject in its own right. Given the significance of the dermal route, precise determinations of percutaneous absorption are key components of the overall assessment of the absorbed dose of the pesticide for risk assessment.

52.2.2  Exposure by Inhalation Setting aside volatile pesticides for the moment, the only spray droplets or particles that pose a potential risk comprise the so-called inhalable or inspirable fraction, which is the mass fraction of airborne particulate capable of entering the respiratory tract via the nose and the mouth, so ­ providing 1127

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a source of absorption into the body, either from direct inhalation or from subsequent oral absorption. This is considered to be the most important indicator of potential inhal­ ation exposure (ACGIH, 1985; Vincent and Mark, 1987). The inhalable fraction depends on the speed and direction of the air movement, on the rate of breathing, and on other factors. For sampling purposes, inhalable particles can be considered to have a mass median diameter of 100 m or less. The respirable fraction is the mass fraction of inhaled particles, which penetrates to the unciliated airways. For sampling purposes, respirable particles can be considered to have a diameter between 0 and 15 m (ISO, 1995). In risk assessment it is common to assume that volatile airborne pesticides are completely retained and absorbed via the respiratory tract, unless there are specific data to demonstrate otherwise. Inhalation exposure is usually a small fraction of the total exposure and can, in some cases, be ignored, for example, the mixing and loading of liquid formulations, particularly if a closed loading system is involved. Conditions under which exposure by the inhalation route becomes important usually involve the use of volatile pesticides or of dusts, fumigants, and sprays, especially in enclosed spaces. It should, however, be borne in mind that a higher proportion (up to 75%; Ross et al., 2001) of the inhaled dose may be retained systemically, compared with the proportion absorbed after dermal exposure, which could be as low as 1% or less of the available dermal dose.

52.2.3  Oral Exposure Some of the larger airborne particulates may be trapped in the mouth or nasal passages and subjected to oral ingestion. Some of the exposure, which is measured as inhal­ ation, might indeed be trapped and absorbed in this way. No serious attempts have been made to measure separately the amount of exposure by this route because of the obvious difficulties involved. Biological monitoring takes into account all routes of absorption, but it is usually unable to distinguish between their relative contributions.

52.3  Previous reviews and guidance on methodology Durham and Wolfe (1962), Wolfe (1976), and Davis (1980) were responsible for the earliest reviews of methodology. These reviews and particularly the methodology of Durham and Wolfe (1962) were used to develop the World Health Organization (WHO) standard protocol for the measurement of exposure (WHO, 1975, 1982). The first WHO protocol (1975) advocated the Durham and Wolfe “patch” method to estimate dermal exposure and included a reference to biological monitoring through the use of

Hayes’ Handbook of Pesticide Toxicology

cholinesterase activity measurement for organophosphorus insecticides. The revised protocol of 1982 proposed an alternative method for the measurement of dermal expos­ ure, the “whole body” sampling technique, and gave an overview of biological monitoring as a means of measuring absorption arising from all routes of exposure. This revised protocol was used to develop the U.S. National Agricultural Chemicals Association (NACA) guidelines for mixer-loader-applicator exposure studies (Mull and McCarthy, 1986). These guidelines placed primary emphasis on the use of the Durham and Wolfe (1962) patch methodology rather than the whole body technique. NACA also published guidelines for conducting field biological monitoring studies as a means of measuring the absorption of pesticides (NACA, 1985). Both the U.S. Environmental Protection Agency (U.S. EPA, 1987) and NACA guidelines contained detailed reviews of the advantages and limitations of the methods for the measurement of exposure to and absorption of pesticides. The International Group of National Associations of Manufacturers of Agrochemical Products (GIFAP), now known as the Global Crop Protection Federation, published a position paper on general aspects of monitoring studies for the assessment of worker exposure to pesticides (GIFAP, 1990). These guidelines were intended to inform the nonspecialist of the various approaches to exposure/ absorption evaluation and their significance. Curry and Iyengar (1992) reviewed and compared the currently available guidelines (published and unpublished) for the evaluation of exposure to individuals using pesticides and those exposed to residues in indoor and outdoor environments. Harmonized exposure and biological monitoring guidelines were proposed in a workshop held in the Netherlands (Chester, 1993; Henderson et al., 1993). The guidelines agreed on at this workshop were further discussed at a Health Canada/North Atlantic Treaty Organizationsponsored Workshop on Methods of Pesticide Exposure Assessment held in Canada in 1993, resulting in a draft guidance document to be submitted to the Organization for Economic Co-operation and Development (OECD). A designated peer review group established at the workshop revised the guidance document, which was submitted to the OECD and published as an OECD Guidance Document (OECD, 1997). The U.S. EPA meanwhile published a series of test guidelines in 1996 describing their preferred approaches to passive dosimetry and biological monitoring for exposure during indoor and outdoor occupational and residential use of pesticides (U.S. EPA, 1996). The guidance in these OECD and U.S. EPA documents represents the most up-todate, harmonized approaches to the assessment of exposure to pesticides, and the reader is referred to them for more detailed information.

Chapter | 52  Worker Exposure: Methods and Techniques

52.4  Design of agricultural worker exposure studies The purpose of a worker exposure and/or biologic­al monitoring study is to generate data for use in a risk assessment. Good study design is therefore a key consideration in ensuring that relevant and useful exposure data are obtained. In deciding on the basic methodological approach, reference can be made to the tiered approach to exposure and risk evaluation to determine whether passive dosimetry will suffice or whether the use of biological monitoring is warranted to give the most accurate determination of the dose absorbed by the worker (Henderson et al., 1993). Agricultural worker exposure studies can be regarded as being of two types: pre- and re-registration studies and postregistration surveillance studies (OECD, 1997). Studies of the first type involve the test subjects complying fully with the requirements of the product label, in particular, use of protective clothing and equipment, application rates, and cleanup procedures. By ensuring compliance, certain constraints are imposed on the activities of the workers that may influence the amount and variability of their exposures. Studies done according to these criteria are also appropriate for inclusion in generic exposure databases since they will have been conducted according to a set of standardized principles. Studies of the second type are done primarily in support of product stewardship and postregistration evaluation of actual pesticide use conditions and practices. Adverse effects on the health of workers might be reported, or there might be a possible need to study the extent of compliance with product label precautions and recommendations. Therefore, the study design should take into account the need to measure exposure under the actual conditions of use and would be free of the constraints imposed on pre- or re-registration studies. Other factors that may influence the sampling strategy include possible concern about specific work activities during pesticide use, including the use of nonstandard application equipment. Passive dosimetry enables separate measurements of the respective contributions of these activities to the total exposure. Identification of the differences in the magnitude of exposure attributable to these activities permits the use of different regulatory proposals for reduction of exposure to acceptable levels. An example is the recommended use of additional protective equipment during procedures with greater potential for exposure, for example, handling and mixing of the concentrated formulation. In designing a study, consideration should be given to whether passive dosimetry and biological monitoring should be conducted concurrently. Both may be justified to provide data for inclusion in generic databases, to examine the relationship between exposure and absorption, and to provide a second measurement should one fail. The U.S.

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EPA, in their latest test guidelines, however, is strongly opposed to this idea (U.S. EPA, 1996). Their concern stems from the reasonable perception that the dermal passive dosimeters, in particular, intercept pesticide as it comes into contact with the worker, thus interfering with the process of dermal contamination and absorption and reducing the biological monitoring. However, if clothing dosimeters are representative of what workers normally use, then this disadvantage can be overcome. This matter is discussed further in Section 52.6. Concerning the number of measurements of expos­ ure and/or absorbed dose in the field study, as a general guide, OECD (1997) proposes a minimum of 10 different subjects. The U.S. EPA (1996) requires a minimum of 15 replicated measurements of exposure, not necessarily in different test subjects. Factors to consider in making the decision on the number of subjects are the following: l l l l

The likely end use of the data The nature of any identified toxicological endpoint The required level of statistical confidence The overall manageability of the study

Where feasible, subjects should be randomly selected from the worker population. It is recommended that a sufficient number of measurements be made in different locations to cover the range of use procedures, conditions, and application equipment for which exposure data are required. Variability in exposure can be addressed by increasing the number of subjects rather than repeatedly monitoring the same individuals. Variability between workers is typically greater than that within the same worker (Kromhout et al., 1993; Rappaport, 1991). In addition, as many sites as possible should be included rather than having subjects use the same equipment under the same conditions. This applies particularly if location is believed to have a significant impact on the variability of the exposure measurements. Certain types of pesticide application procedures render management of the study difficult, such as those involving aerial application. In such cases, repeated monitoring of the same individuals is a possible option, although the limitations of such a choice should be recognized. Should biological monitoring be necessary, a further limitation is introduced in that interindividual variation in metabolism of the pesticide would be less well evaluated. The duration of the study will be prolonged owing to the possible need to collect urine samples over several days for each individual. The consequence is that repeat monitoring cannot commence until urine collection is complete. These difficulties should be considered when using biological monitoring in such circumstances. Ideally, the duration of a single measurement of expos­ ure or absorbed dose should be representative of the typical working day, so that all the work activities that contribute

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to total exposure, such as equipment repair and cleanup, are assessed. This criterion applies particularly to studies involving biological monitoring in which the skin is the predominant, if not only, route of exposure and absorption. For many pesticides that are not well absorbed percutaneously, absorbed dose data should not be linearly extrapolated on the basis of time or amount of active ingredient used in the same way as passive dosimetry exposure data. The percutaneous absorption of a pesticide depends on the rate at which it is absorbed, the area of skin contaminated, and the duration of skin contact. The rate will increase up to the maximum steady-state rate. Once this is achieved, further increases in deposition of pesticide on the skin will have no further impact on the absorption process, as it is saturated. Certainly, an increase in deposition can result in increased absorption up to maximum rate, but not in direct proportion (Chester, 1988). In studies involving passive dosimetry of volatile or unstable pesticides, a shorter monitoring period should be considered. This would be based on a consideration of the physical-chemical properties of the compound. The choice of monitoring duration should account for the possibility of dosimeter saturation. Ideally, a single set of dosimeters should be used per worker; however, a change of dosimeters during the workday may be necessary if different tasks are to be monitored separately. This can, however, be difficult to manage from the standpoint of practicality. The choice of use pattern (including application equipment) should account for factors such as whether it is the predominant one for the product or a minor one for which no generic data are available to the investigator or regulatory authority to enable a risk assessment. In pre- and reregistration studies, the product should be used in the study at a representative, recommended rate of application and on the likely maximum area of crop treatable in a working day under local conditions. It should also be applied in accordance with all the label recommendations for use. These principles also apply to studies involving postapplication crop re-entry in which exposure should be determined after the shortest permissible re-entry period, if known. In postregistration surveillance studies, these criteria should not be enforced; this ensures maximum representativeness of actual use conditions and exposure variability. Where the product label recommends the use of protective clothing and/or equipment, these items should be provided to the subjects in pre- and re-registration studies by the supervisory team to ensure standardization. This will benefit the scientific interpretation of the data, because the variable of differing standards of protection by different types and conditions of protective equipment will have been removed. Inclusion of criteria such as these ensures exposure and risk assessment for the product in accordance with the label recommendations for regulatory use. However, efforts should be made to ensure that the recommended protective clothing and equipment are practical

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and realistic to use under local conditions. In postregistration surveillance studies, the study team should not mandate use of protective clothing and equipment, although ethical or legal viewpoints on product label recommendations in this respect should be considered.

52.5  Test subjects Agricultural workers should be the test subjects in a field study, rather than inexperienced volunteers. If this is not possible, use of nonprofessional personnel may have to be considered, provided that they are given the requisite training in the handling and use of the pesticide and equipment. The disadvantage of this choice is that the subjects would not necessarily be representative of the worker population. Males or females may be considered for inclusion in passive dosimetry studies. However, in studies involving biological monitoring, the decision is more difficult owing to the potential impact on the interpretation of the metabolite excretion data due to possible sex differences in metabolism and kinetics. If it is likely that the product will be used by both sexes then it is important to know whether there are differences. All subjects should be asked to provide written, informed consent to participate in a study after they are provided with the requisite information on the pesticide. Potential subjects must be informed that they are free to withdraw from the study at any time. Subjects should be screened for any pre-existing medical conditions that may be affected by use of the pesticide, depending upon its toxicological profile. Depending upon the circumstances and local custom, it may be appropriate to provide the subjects with information on their individual results.

52.6  Methods for measuring exposure (passive dosimetry) 52.6.1  Patch Method for Dermal Exposure In this method, the potential contamination of the workers’ skin and clothing is measured using a variable number of absorbent cloth or paper patches attached to body regions inside and outside clothing. The surface area covered by the patches represents less than 10% of the total body surface area. After a defined or measured period of exposure, the patches are removed and analyzed for pesticide content. The quantity of a pesticide on a patch of known area is then related to the area of limb or other body part on the assumption that deposition is uniform over the body parts. Body part surface areas can be obtained from standard reference texts and exposure guidance documents, the most recent of which are those of the U.S. EPA (1996, 1999) and the OECD (1997). The assumption of uniform deposition is perhaps the principal disadvantage of the patch technique. This is

Chapter | 52  Worker Exposure: Methods and Techniques

i­llustrated by the extrapolation of the value given by half the limit of quantification to the total body part; this may give a substantial under- or overestimate of exposure. The principal disadvantage can be mitigated to a certain extent by increasing the number of patches located on body parts likely to receive significant exposure. Individual body part exposure values are then added to give a total potential exposure expressed in mg/h, mg/day, or mg/kg product handled or applied. In the WHO protocol fewer patches are recommended, representing only 3% of the body area. In both cases standard body part surface areas are used in correcting individual patch values (WHO, 1982; U.S. EPA, 1996). According to the WHO protocol, only the pesticide contacting the normally unclothed area of skin, for example, head, neck, hands, and forearms, is used to calculate actual exposure. In temperate climates about 10% of the total body area is normally unprotected during use of pesticides and other agricultural activities. Normal work clothing, such as cotton trousers and shirts, is absorbent and may retain and allow penetration of a proportion of the pesticide contamination (Driver et al., 2007). Therefore, it is still necessary to estimate exposure to the covered areas of the body. Consequently, the amount of pesticide penetration is often measured using patches attached beneath the clothing. Without such measurements, an estimate of penetration of normal clothing must be used. However, penetration of clothing by pesticides is a highly variable process, which is influenced by factors such as the type of formulation (liquid or solid) and the amount or volume of deposition on the clothing, dampness of the clothing, pesticide vapor pressure, the location of deposition (e.g., seams), and the type of fabric. Further uncertainties are introduced by the method of sampling for clothing penetration. The patch method may give significant under- or overestimates of exposure, depending on whether the patches have captured the nonuniform, random deposition of concentrate splashes or spray droplets. This limitation applies equally to crop re-entry procedures, such as harvesting, where contact with the crop is not a uniform process. Despite the readily apparent limitations, the patch method remains a useful method for exposure evaluation.

52.6.2  Use of Fluorescent Tracers and Visible Dyes: Quantification by Analysis or Video Imaging Dermal exposure can be quantified directly by measuring deposition of fluorescent materials or visible dyes on the clothing and/or skin. The fluorescent tracer or dye can be substituted and extracted from passive dosimeters and analyzed in the same way as the pesticide. By adjustment for concentration differences, an estimate of exposure to the pesticide can be obtained, similar to extrapolation from

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one monitored pesticide to another used in the same way. A video imaging/fluorescent tracer technique has been developed (Fenske et al., 1986a, b; Fenske, 1990). This method involves the incorporation of a fluorescent tracer in a pesticide formulation and subsequent visual and quantitative analysis using a video imaging technique. It reveals nonuniform patterns of exposure that escape detection by the patch method. It also demonstrates that exposure can occur beneath protective clothing. An important advantage of this technique is that the skin serves as a collection medium rather than dosimeter patches or clothing. The main limitations of this method are the assumptions that the relative transfer of the tracer and the pesticide in the field and their permeation of the clothing are equivalent. Another acknowledged limitation is the limited dynamic range of in situ fluorescence. Specifically, high exposure areas, for example, hands, may accumulate concentrations that cannot be accurately measured due to self-quenching. However, these assumptions are analogous to those involving use of generic exposure databases. That is, the expos­ ure to a pesticide measured under a given set of conditions is assumed to represent the exposure associated with a second pesticide under the same conditions. Ancillary studies can assess possible differences in the relative transfer of the tracer and the pesticide. The techniques are particularly useful for the training of operators by demonstrating the extent of their contamination, thus enabling modification of their working practices to reduce exposure. Roff (1994) developed the technique a stage further by using a dodecahedral lighting system to illuminate the contaminated worker. Measurement errors inherent in the Fenske technique caused by body surface morphology are apparently reduced. This rather large piece of equipment has been used under field conditions If a tracer compound or visible dye is chosen, its performance and suitability as a surrogate should be validated before the field study. Apart from the usual criteria of quality control acceptability applicable to all pesticides, the surrogate compound should not significantly alter the physical properties of the formulation or spray mixture. The key question that determines the utility of a tracer or dye is whether it affects, is retained by, or penetrates clothing similar to or in parallel with the pesticide of interest. The technique can only be used in exposure studies and not in biological monitoring (although it can be used with concurrent biological monitoring of a pesticide). Its greatest utility probably lies in substituting for pesticides that are particularly unstable during the sampling and analytical phases.

52.6.3  Whole Body Method The whole body method came into use during the late 1970s/early 1980s (Abbott et al., 1987; WHO, 1982). The method involves the use of clothing, usually two layers of

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cotton or cotton/blend material that act as the pesticide collection media. The outer layer of clothing should be representative of what the workers might wear under normal circumstances. The inner layer, usually a set of combination “long johns,” represents the skin. This method overcame one of the inherent problems of the patch method, that is, the assumption of uniformity of pesticide depos­ ition on the skin and clothing. Any additional protective clothing and equipment recommended for the product under study are worn over the sampling clothing, thus enabling an evaluation of their protection efficiency. The use of the whole body method overcomes the perceived problem of nonuniformity of deposition. Furthermore, extrapolation from small target areas to larger body regions is not necessary. For these reasons, the method is believed to give a more accurate estimate of potential and actual dermal exposure. The whole body method can be adapted for concurrent use with biological monitoring by use of work clothing as dermal dosimeters, which represents what the workers would normally wear under the prevailing conditions. This is contrary to the view expressed by the U.S. EPA (1996). Whereas the patch and standard whole body methods place sampling media between the pesticide and the clothing or skin, thus acting as a barrier interfering with the normal process of skin contamination and percutaneous absorption – the U.S. EPA’s main concern – the advantage of this method is that the capture, retention, and penetration properties of the normal work clothing are mimicked as closely as possible. It is important, therefore, to have an understanding of the range of normal work clothing worn by the worker population under study. Clothing for sampling should be selected cautiously, using the minimum that might be worn under the prevailing conditions. Therefore, the assessment of residual clothing contamination and transfer to the skin beneath the clothing is as realistic as possible. This method is particularly relevant for North European and North American temperate countries where the typical work clothing consists of a t-shirt, long-sleeved shirt, socks, and long trousers, and/or coveralls. Actual exposure of the skin beneath the clothing can be estimated by determining the ratio of outer to inner clothing penetration or transfer of the pesticide. An obvious limitation is that the permeation and transfer properties of the outer and inner clothing are assumed to be the same. For analytical considerations, it may be necessary to use noncolored, white materials such as cotton or cotton/polyester mixtures. As in the standard whole body method, the clothing is sectioned into individual body parts and analyzed separately to determine the regional distribution of total potential and actual exposure. Few attempts have been made to validate methods for the monitoring of dermal exposure, for example, by using biological monitoring to compare the derivations of absorbed dose. Ross et al. (2008) undertook a comprehensive evaluation of studies that involved both passive

Hayes’ Handbook of Pesticide Toxicology

d­ osimetry (PD) and biological monitoring (BM) to determine the validity of the PD methods on the basis that BM represents the “gold standard” measurement of pesticide absorbed dose. A total of 14 concurrent or consecutive PD:BM studies were evaluated, with 18 different methods of application or re-entry scenarios for eight different pesticides for which measured human kinetics and dermal absorption data were available. The evaluation demonstrated that the total absorbed dose estimated using PD was generally similar to the measurements for the same scen­ arios using BM. The statistical analysis of PD:BM ratios for individual workers showed them to be significantly correlated, thereby demonstrating the validity of the PD methodologies. Despite this comprehensive evaluation, all methods should be viewed as providing only an approximate estimate of the dermal exposure. Dermal exposure methods will still be needed, as biological monitoring cannot be applied to all pesticides. Exposure method development is therefore a continuing need.

52.6.4  Hand (and Head) Exposure Measurement of hand exposure is one of the most important aspects of a study to monitor worker exposure. The contribution of the hands to total exposure is well documented and was originally recognized in the seminal works of Batchelor and Walker (1954) and Durham and Wolfe (1962). The U.S. EPA (1996) reviewed the literature on studies that had included hand exposure measurements and concluded that its contribution to total exposure ranged from around 40 to 98%, depending upon the application method. Dermal ex­posure incurred during handling pesticide formulations, for example, during mixing, loading, and preparation for application, predominantly involves the hands, depending upon the type of equipment that is being loaded. Head exposure is usually negligible in relation to other body parts, representing typically 1% of total dermal exposure. Exposure is usually measured with a dosimeter such as a patch on the forehead or a cap. Alternatively, a face and neck wipe or wash technique is used, adapted from that developed for the hand measurement. Therefore, the following discussion that pertains to hand wiping or washing can be considered relevant also for face and neck wiping or washing. The methods for measuring hand exposure include using lightweight absorbent gloves or sections cut from gloves and wiping or rinsing the hands in various solvents, for example, 95% ethanol (U.S. EPA, 1996). Mild detergent solutions can be used in the hand wash technique, for example, Aerosol OT. An important distinction must be drawn between the solvent or detergent hand rinse and the hand wash techniques. Rinsing usually involves inserting one or both hands in a polyethylene bag containing the solvent or detergent followed by a relatively mild movement or shaking of the hands to remove the contaminant.

Chapter | 52  Worker Exposure: Methods and Techniques

Washing is rather more vigorous than rinsing and usually involves application of diluted detergent to the hands of the worker followed by vigorous washing of the hands together over a collection bowl. Water or the diluted detergent is then poured over the hands with collection of the removed contaminant in the bowl. Rinsing the hands with a solvent such as ethanol might cause skin damage or disrupt skin barrier function and enhance percutaneous absorption of the pesticide. However, in the many studies that have used this technique there has been little evidence that these effects occurred. The generally held view is that the use of gloves results in a significant overestimation of hand dermal exposure, owing to the retention of more of the pesticide on a multi­ layer porous medium than would otherwise be retained by the skin surface of the hands. However, there might be circumstances in which gloves underestimate the potential exposure of the hands, for example, if breakthrough occurs due to glove saturation, leading to material being deposited on the skin of the hands and not retained by the gloves. Gloves also contain foreign materials such as sizing, which may be co-extracted with the pesticide. At low levels of contamination this may cause analytical difficulties. However, glove contamination with dirt and grease arising from the worker’s activities are a more likely cause of analytical problems. All these methods have advantages and limitations, and it is difficult to evaluate the accuracy of any procedure, particularly when there is no real “gold standard” measurement that has direct relevance to the field environment. This concern about measurement accuracy and lack of standardization in measurement technique has created uncertainty over the validity of the early generic hand exposure data in databases such as the Pesticide Handlers Exposure Database (PHED) in North America and the European Predictive Operator Exposure Model (EUROPOEM) in Europe (AIR, 1996; Versar, 1991). This has been addressed to a large extent by the industry task forces – Agricultural Re-entry Task Force (ARTF) and Agricultural Handlers Exposure Task Force (AHETF) in the United States and the European Crop Protection Association (ECPA) Operator Exposure Monitoring Task Force (EOEM TF) – adopting basically the same hand wash method involving the diluted detergent Aerosol OT for their exposure studies comprising their respective databases from the mid-1990s onward. Despite this standardization, there has been a long-standing reservation, particularly within the academic and regulatory communities, about the ability of the rinse and wash techniques to accurately determine hand exposure. Specifically, there is a belief that these techniques might underestimate hand exposure, leading to systematic bias. The perceived underestimation derives from possible inefficient recovery from the skin of the hands and percutaneous absorption of the pesticide. The first of these reasons is potentially the more significant for the vast majority of pesticides because

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percutaneous absorption tends to be only a small fraction of the applied dose, with an average of 10% for pesticides applied to human skin (Ross et al., 2001). Two compounds radically influenced this average (orthophenyl phenol and propoxur). Davis et al. (1983) were the first researchers to publish results of a comparison of two methods for measuring hand exposure. The study involved apple thinners working in orchards treated with azinphos-methyl. A comparison was made between alcohol hand rinses and gloves (cotton and nylon) showing that measurements with gloves were approximately fivefold higher than those with rinses. There was no attempt to estimate the true hand exposure so that both underestimation and overestimation were possible with the rinse and gloves methods, respectively. Fenske et al. (1989) demonstrated the influence of exposure duration on the differences in hand exposure measurements with hand rinses and cotton gloves in peach harvesters in orchards treated with captan. The absorptive capacity of the gloves was greater for short exposure durations of 0.5 and 1 h than for longer durations of 1.5 and 3 h. The gloves gave significantly higher measurements than hand rinses at the shorter durations. The measurements were not significantly different at the longer durations. Glove loading may have been the explanation for the lack of significant differences for the longer durations. Decreases in the absorptive capacity of the gloves may have been due to moisture, soil, and/or sweat combined with captan residues. Breakthrough beneath the gloves was also noted. These technical issues and uncertainties underpinned the idea that use of such techniques in a worker exposure study should be preceded by a laboratory-based validation of their sampling efficiencies, possibly using human subjects. Fenske and Lu (1994) investigated the recovery efficiency of a solvent hand rinse method for the insecticide chlorpyrifos. Their findings suggest that exposure to pesticides such as chlorpyrifos that are well adsorbed to the skin cannot be estimated accurately by the hand rinse method. However, Geno et al. (1996) showed for chlorpyrifos and pyrethrin that hand wipes with isopropanol removed in excess of 90% of the material applied to hands. Fenske et al. (1999) compared the hand exposures of orchard apple thinners to azinphos-methyl using three methods: glove, hand wash, and wipe. Hand exposure estimates derived from the three methods differed significantly. Based upon the hand wash measurements and the laboratory recovery/efficiency study, it was concluded that the glove method gave a 2.4-fold overestimate whereas the wipe method gave a 10-fold underestimate. The authors also concluded that methods should be validated and standardized to enable the development of more accurate hand exposure estimates. Brouwer et al. (2000) compared removal efficiency results from different hand rinse and wash procedures for various pesticides and one nonpesticide. The studies done by Brouwer and colleague used direct “spiking” of material

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onto the hands, whereas Fenske et al. had used subjects hand­ ling and contacting spiked test tubes. The rinse efficiencies ranged from 22 to 96% with a median of 73%. An additional finding was that wipe sampling is less efficient than the rinse technique. The authors noted that identified sampling protocols, including sampling techniques, differed with respect to key technical issues. They recommended that sampling efficiency studies should be done before field sampling, under conditions similar to those likely to be encountered in the exposure process in the field, for example, levels of skin loading and time of residence of the compound on the skin. In studies involving concurrent passive dosimetry and biological monitoring, a hand-washing procedure involving standardized detergent and water can be used. This procedure is identical to that described earlier as the hand wash method, except that the measurement is taken only when the workers would normally wash their hands. The basis is that when the total absorbed dose is determined with biological monitoring there should be no interference with the normal process of dermal contamination and percutaneous absorption. Therefore, the use of gloves or a solvent washing or rinsing technique is inappropriate, because these methods would retain pesticide otherwise available for absorption or potentially disrupt the barrier function of the skin. Inevitably, there is a loss of standardization of the intervals at which samples are taken. However, it does give some information on the extent of hand exposure that might be of value in overall data interpretation. In 2007, the U.S. EPA convened a Scientific Advisory Panel (SAP) meeting to review several aspects of expos­ ure measurement and assessment methodologies, in direct response to concerns expressed about the proposed study designs and protocols planned by the AHETF and AEATF (Antimicrobial Exposure Assessment Task Force) for new studies to provide exposure data for their respective generic exposure databases (U.S. EPA, 2007). Among the various topics addressed were the hand exposure methods and, in particular, whether a correction should be applied for hand wash recovery efficiency or to compensate for material absorbed percutaneously during sample collection periods. There was no distinction drawn between the rinse technique and the more vigorous wash technique, perhaps because of a lack of supporting data. The panel was slightly equivocal about a need to correct hand wash data for the efficiency at recovering pesticides from the skin. The main concern was with the unknown and variable recovery efficiencies reported in the literature. A first-order kinetic model of adsorption that depends upon the pesticide’s KO/W or octanol-water partition coefficient was suggested. Limited results using this model indicate that the interval from initiating exposure until washing the hands can be important when measuring pesticide handler expos­ ure over the planned 4- to 8-h day. However, some panel members pointed out that the accuracies of either modeling or experimental data could be confounded in field

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conditions by the effects of repetitive (multiple) rinsing or washing that can change the skin’s absorption rate, either enhancing or decreasing recovery from the skin. Overall, the panel recommended that use of a hand-washing technique should be accepted in AHETF or AEATF studies if it is supported by either laboratory data and/or a model that predicts and can correct for its efficiency over the sampling time for the pesticide being studied. The choice of hand exposure measurement technique is really between the hand wash with either mild detergent and water or a solvent compatible with skin having high capacity for dissolving the pesticide of interest, and the glove. The wipe and rinse methods with solvent or detergent may be too passive and lacking in vigor. The hand wash with detergent and water, simulating exactly how workers wash their hands in reality, would seem to have this essential vigor, particularly if underpinned by data indicating efficient removal of the compound from the skin surface, for example, from percutaneous absorption studies in which the skin is gently washed after an 8-h or less contact period. If these data show efficient removal of material with a mild detergent and water wash as per recommended protocol, this should provide sufficient confidence to use the wash technique in the field study. The conduct of a laboratory study involving application of a pesticide to the hands of human volunteers should be undertaken only as a last resort. Apart from the ethical considerations and associated complications of a human study in the current climate to evaluate the efficiency of removal of a pesticide from a single body part, there is the added problem of simulating accurately the physical–chemical matrix of the pesticide mixed with the cocktail of extraneous substances such as dust and grease on the workers’ hands. Ross et al. (2008) demonstrated that pesticide absorbed doses determined with passive dosimetry were statistically indistinguishable from the “gold standard” biological monitored absorbed doses. The studies evaluated in this comparison all involved the detergent/water hand wash. The fact that the doses were similar in the two methodological approaches indicates that the hand wash did not underestimate hand exposure and was sufficiently accurate. Given that the fundamental need is for an accurate determination of hand exposure, that is, the bioavailable amount of pesticide in contact with the skin from which percutaneous absorption can occur under actual conditions of pesticide use, the hand wash is considered to meet this need better than the alternatives and is recommended as the method of first choice.

52.6.5  Inhalation Exposure Exposure by inhalation is usually a minor route of absorption in comparison with the dermal route. There are exceptions to this, for example, when dusts, fine aerosols, and

Chapter | 52  Worker Exposure: Methods and Techniques

fumigants are applied or when materials are applied indoors. The U.S. EPA (1996) reviewed several exposure studies and found that the inhalation route contributed negligible amounts (nondetectable) to about 9% of total exposure. In most cases, the contribution was less than 1%. The extent of the contribution depends upon the method of application, whether used outdoors or indoors and on factors such as the volatility of the pesticide. Significantly, for pesticides that are poorly absorbed via the skin, the inhalation route can become the most important route of absorption. Van Dyk and Visweswariah (1975) and Lewis (1976) reviewed the methodology for field monitoring of airborne pesticides. The former reviewed the sampling media available for collection of pesticides but with particular emphasis on static environmental sampling rather than personal sampling. A personal air sampling method is the most appropriate for the determination of potential inhalation exposure of workers. Several techniques are available such as gauze pads in place of filters in respirators for agricultural use, pioneered by Durham and Wolfe (1962), midget impingers, solid adsorbents, and filter cassettes attached to batterypowered personal sampling pumps. A personal sampling technique involving sampling devices located in the breathing zone and sampling pumps is preferred for reasons of practicability and representativeness. Breathing rates for the calculation of inhalation exposure from airborne concentration data can be obtained from standard reference texts such as the U.S. EPA’s Exposure Factors Handbook (1999). The advantage of using the modified respirator is that the subject produces the airflow so that breathing rate and total volume of air inhaled do not have to be estimated. However, the gauze pads must be capable of trapping the pesticide efficiently. The respirator must also fit properly to the face. Perhaps the main disadvantage of this technique is that the subject must wear the respirator for the duration of the monitoring period, which ideally should be a complete working day, which may cause discomfort. Midget impingers, which traditionally have used ethy­l­ ene glycol as the trapping medium, have a long history of use in measuring agricultural worker inhalation expos­ure. The technique suffers from the major drawbacks of spillage of the trapping liquid and inefficient trapping and retention of some pesticides. Micro-impingers have been developed, which overcome the first of these drawbacks. Personal air samplers allow the use of respirators or dust masks for protection, if required by the product label. However, they do not measure the true exposure of workers wearing respiratory protection. The choice of sampling medium is determined by the nature of the pesticide. A filter cassette or sampling head should be used for spray particulates and a solid adsorbent material for volatile compounds. The inhalable fraction (all material capable of being drawn into the nose and mouth) is the most biologically

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relevant fraction to measure. An example of a suitable device is the Institute of Occupational Medicine personal sampling head designed specifically to collect this fraction (Vincent and Mark, 1987). For use of this device, a sampling flow rate of 2 l/min is a specific requirement. There are now commercially available devices manufactured by SKC and TSI that collect, as they separate fractions, the inhalable and respirable components. Examples of suitable adsorbent materials for some vola­ tile compounds are activated charcoal, Tenax, and XAD-2 resins mounted in stainless steel or glass tubes. The choice of material should be determined by analytical retention (trapping efficiency) and extractability studies. Concurrent sampling for particulates and vapor can be achieved by mounting the filter sampling head in front of the vapor trap in a “sampling train.” This train allows retention on the resin of any vapor stripped off the filter. The material on the filter can be analyzed gravimetrically and/or chemically and an estimate made of the pesticide content of the particulate sample. If use of such a sampling train is needed, laboratory validation of the sampling efficacy, particularly of the adsorbent resin, is necessary owing to the possibility of stripping material from the resin by the relatively high flow rate of 2 l/min. The measurement of the inhalable fraction with use of adsorbent resins for the vapor phase is recommended as the method of first choice.

52.7  Methods to measure the absorbed dose 52.7.1  Biological Monitoring Biological monitoring of pesticide workers was first used as a means of assessing health effects or modification of biochemical parameters as a consequence of exposure to organophosphorus compounds by measurement of plasma cholinesterase levels (for example, Peoples and Knaak, 1982). This type of assessment, used in the chemical industry for many years, can be termed biological effect monitoring and must be distinguished from the type of monitoring that determines the absorption of chemicals by measuring the chemical or its metabolites in body fluids, usually urine, blood, or exhaled breath. Analysis of body fluids and excreta, usually urine, for parent compound or metabolites can provide both a qualitative and a quantitative measurement of absorbed dose for pesticides that lend themselves to this form of monitoring. The technique has a distinct advantage over passive dosimetry because it evaluates actual, rather than potential, absorption. It integrates absorption from all routes of exposure: dermal, inhalation, and primary and secondary oral ingestion. However, it is difficult to differentiate the contributions to the absorbed dose from different aspects of the work procedures or to distinguish between the

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relative contributions of the different routes of exposure to the total absorbed dose. Van Heemstra-Lequin and Van Sittert (1986), Wang et al. (1989), and Henderson et al. (1993) reviewed biological monitoring in the context of pesticides. The OECD (1997) provided detailed guidance on how to conduct biological monitoring studies. Early biological monitoring studies on pesticides were able to demonstrate absorption without quantifying the amount of absorption. For example, Durham and Wolfe (1962) measured p-nitrophenol, a metabolite of parathion, in the urine of workers; Swan (1969) measured paraquat in the urine of backpack spray operators. The absorbed dose of a pesticide can only be quantified accurately if the metabolism and pharmacokinetics of the compound are understood, ideally from human studies. Woollen (1993) reviewed the specific requirements for this type of biological monitoring study. Studies that have met these requirements include those on the herbicide fluazifopbutyl (Chester and Hart, 1986) and 2,4-dichlorophenoxyacetic acid amine (Grover et al., 1986; Ritter and Franklin, 1989). Data from human dosing studies facilitate the design of a field sampling strategy and secondly define the body fluid matrix of choice. Urine is the matrix of choice as its collection is noninvasive and the collection of 24-h urine samples is practicable. Complete 24-h urine collections are desirable, and this can be checked in a number of ways, for example, by measuring the concentration of creatinine. Substantially incomplete collections are readily apparent, and these samples are either excluded or an allowance is made, for example, by use of a correction factor based upon the average daily urine volume for the individual concerned (Woollen, 1993). Specific gravity and osmalarity are alternative means of checking for completeness of urine collection (Alessio et al., 1985) Unstable or highly volatile pesticides are not good candidates for passive dosimetry, despite the efforts to accurately assess field, storage, and transit losses. Biological monitoring should be considered for these pesticides and may be the only means of obtaining adequate quantitative data from which the absorbed dose can be derived. An example of the successful use of biological monitoring to estimate exposure occurring primarily by the inhalation route is urinary monitoring for a metabolite of the fumigant 1,3-dichloropropene (Osterloh et al., 1989; Van Welie et al., 1991). If a human metabolism study were impracticable, then animal metabolism data might be used; if metabolism and excretion kinetics are similar in several animal species, then it could be assumed that humans will metabolize and excrete the compound in a similar manner. This carries a degree of uncertainty. There are examples, described in detail by Woollen (1993), that demonstrate the limitations of this approach. Apart from interspecies differences in metabolism, there is the possibility of dose-dependent differences, which might necessitate metabolism studies

in animals and humans at doses similar to the anticipated worker exposures. Pesticides that are extensively metabolized to a large number of metabolites are not good candidates for biologic­al monitoring. The absorbed dose cannot be determined accurately using data on a minor metabolite, particularly if there is wide interindividual variation in the proportion of the parent compound excreted as this metabolite. However, a minor metabolite might provide some useful information as a biological indicator in the absence of more abundant metabolites. Overall, it can be concluded that if the requisite human metabolism data are available for a pesticide, biological monitoring provides the most accurate means of estimating the absorbed dose for quantitative risk assessment.

Conclusion Exposure measurements are integral to risk assessment for agricultural pesticide handlers and postapplication crop reentry workers. Study design is a key consideration, and a clear distinction is drawn between studies done for pre- or re-registration and those for postregistration surveillance of exposure. The methods for the determination of exposure (passive dosimetry) and the systemic, absorbed dose (biological monitoring) are described, with emphasis on their strengths and limitations. Patch and whole body dosimetry and use of fluorescent tracers can all be used to determine dermal exposure, depending on the study objectives and available resources. Personal air sampling methods are used for determining inhalation exposure with collection of the inhalable fraction of airborne particulates and spray droplets, and volatile component, if relevant. Biological monitoring is the “gold standard” method for determining the integrated, absorbed dose arising from all routes of exposure.

References Abbott, I. M., Bonsall, J. L., Chester, G., Hart, T. B., and Turnbull, G. J. (1987). Worker exposure to a herbicide applied with ground sprayers in the United Kingdom. Am. Ind. Hyg. Assoc. J. 48, 167–175. Agriculture and Agro-Industry including Fisheries (AIR). 1996. “The Development, Maintenance, and Dissemination of a European Predictive Operator Exposure Model (EUROPOEM) Database. A EUROPOEM Database and Harmonised Model for Prediction of Operator Exposure to Plant Protection Products.” Draft Final Report, December, Publication No. AIR3 CT93-1370, concerted action under the AIR specific programme of the Community’s Third Framework for Research and Technological Development and managed by DGVI.FII.3. Alessio, L., Bolin, A., Dell’Orto, A., Toffoletto, F., and Ghezzio, I. (1985). Reliability of urinary creatinine as a parameter used to adjust values of urinary biological indicators. Int. Arch. Occup. Environ. Health 55, 99–106.

Chapter | 52

Worker Exposure: Methods and Techniques

American Conference of Government Industrial Hygienists (ACGIH)/ Technical Committee on Air Sampling Procedure (1985). “Particle Size Selective Sampling in the Workplace,” ACGIH, Cincinnati. Batchelor, G. S. and Walker, K. C. (1954). Health hazards involved in the use of parathion in fruit orchards of North Central Washington. Am. Med. Assoc. Arch. Ind. Hyg. 10, 522–529. Brouwer, D. H., Boeniger, M. F., and Van Hemmen, J. (2000). Hand wash and manual skin wipes. Ann. Occup. Hyg. 44(7), 501–510. Chester, G. (1988). Pesticide applicator exposure – towards a predictive model for the assessment of hazard. Aspects Appl. Biol. 18, 331–343. Chester, G. (1993). Evaluation of agricultural worker exposure to, and absorption of pesticides. Ann. Occup. Hyg. 37, 509–523. Chester, G. and Hart, T. B. (1986). Biological monitoring of a herbicide applied through backpack and vehicle sprayers. Toxicol. Lett. 33, 137–149. Curry, P. and Iyengar, S. (1992). Comparison of exposure assessment guidelines for pesticides. Rev. Environ. Contam. Toxicol. 129, 79–93. Davis, J. E. (1980). Minimizing occupational exposure to pesticides: personal monitoring. Residue Rev. 75, 35–50. Davis, J. E., Stevens, E. R., and Staiff, D. C. (1983). Potential exposure of apple thinners to azinphos-methyl and comparison of two methods for assessment of hand exposure. Bull. Environ. Contam. Toxicol. 31, 631–638. Driver, J., Ross, J., Mihlan, G., Lunchick, C., and Landenberger, B. (2007). Derivation of single-layer clothing penetration factors from the pesticide handlers exposure database. Reg. Toxicol. Pharmacol. 49, 125–137. Durham, W. F. and Wolfe, H. T. (1962). Measurement of the exposure of workers to pesticides. Bull. WHO 26, 75–91. Fenske, R. A. (1990). Non-uniform dermal deposition patterns during occupational exposure to pesticides. Arch. Environ. Contam. Toxicol. 19, 332–337. Fenske, R. A., Leffingwell, J. T., and Spear, R. C. (1986a). A video imaging technique for assessing dermal exposure – I. Instrument design and testing. Am. Ind. Hyg. Assoc. J. 47, 764–770. Fenske, R. A., Birnbaum, S. G., Methner, M. M., and Soto, R. (1989). Methods for assessing fieldworker hand exposure to pesticides during peach harvesting. Bull. Environ. Contam. Toxicol. 43, 805–815. Fenske, R. A. and Lu, C. (1994). Determination of handwash removal efficiency; incomplete removal of the pesticide chlorpyrifos from skin by standard handwash techniques. Am. Ind. Hyg. Assoc. J. 55, 425–432. Fenske, R. A., Simcox, N. J., Camp, J. E., and Hines, C. J. (1999). Comparison of three methods for assessment of hand exposure to azinphos-methyl (Guthion) during apple thinning. App. Occup. Environ. Hyg. 14, 618–623. Fenske, R. A., Wong, S. M., Leffingwell, J. T., and Spear, R. C. (1986b). A video imaging technique for assessing dermal exposure – II. Fluorescent tracer testing. Am. Ind. Hyg. Assoc. J. 47, 771–775. Geno, P. W., Camann, D. E., Harding, H. J., Villabos, K., and Lewis, R. G. (1996). Handwipe sampling and analysis procedure for the measurement of dermal contact with pesticides. Arch. Environ. Contam. Toxicol. 30, 132–138. Groupement International des Associations Nationales de Fabricants de Produits Agrochemiques (GIFAP). (1990). Monitoring Studies in the Assessment of Field Worker Exposure to Pesticides. Technical Monograph No. 14. GIFAP, Brussels. Grover, R., Franklin, C. A., Muir, N. I., Cessna, A. J., and Riedel, D. (1986). Dermal exposure and urinary metabolite excretion in farmers repeatedly exposed to 2,4-D amine. Toxicol. Lett. 33, 73–83.

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Henderson, P. Th., Brouwer, D. H., Opdam, J. J. G., Stevenson, H., and Stouten, J. Th. J. (1993). Proceedings of workshop on: risk assessment for worker exposure to agricultural pesticides. Ann. Occup. Hyg. 37, 499–507. International Organization for Standardization (ISO). (1995). Air Quality—Particle Size Fraction Definitions for Health-Related Sampling. ISO 7708:1995(E). Kromhout, H., Symanski, E., and Rappaport, S. M. (1993). A comprehensive evaluation of within and between worker components of occupational exposure to chemical agents. Ann. Occup. Hyg. 37, 253–270. Lewis, R. G. (1976). Sampling and analysis of airborne pesticides. In “Air Pollution from Pesticides and Agricultural Processes” (R. E. Lee Jr., ed.). CRC Press, Cleveland. Mull, R. and McCarthy, J. F. (1986). Guidelines for conducting mixer– loader–applicator studies. Vet. Hum. Toxicol. 28, 328–336. National Agricultural Chemicals Association (NACA) (1985). “Guidelines for Conducting Biological Monitoring – Applicator Exposure Studies,”. NACA, Washington, DC. Osterloh, J. D., Wang, R., Schneider, F., and Maddy, K. (1989). Biological monitoring of dichloropropene: air concentrations, urinary metabolite, and renal enzyme excretion. Arch. Environ. Health 44, 207–213. OECD. (1997). Guidance Document for the Conduct of Studies of Occupational Exposure to Pesticides During Agricultural Application. OECD Environmental Health and Safety Publications Series on Testing and Assessment No. 9. Environment Directorate, OECD Paris. Peoples, S. A., and Knaak, J. (1982). Monitoring pesticide blood cholinesterase and analysing blood and urine for pesticides and their metabolites. In “Pesticide Residues and Exposure” (J. R. Plimmer, ed.), ACS Symposium Series, Vol. 182. Am. Chem. Soc., Washington, DC, pp. 41–57. Rappaport, S. M. (1991). Assessment of long-term exposures to toxic substances in air. Ann. Occup. Hyg. 35, 61–121. Ritter, L., and Franklin, C. A. (1989). Use of biological monitoring in the regulatory process. In “Biological Monitoring for Pesticide Exposure” (R. G. M., Wang, C. A., Franklin, R. C., Honeycutt, and J. C. Reinert, eds.), ACS Symposium Series, Vol. 382. Am. Chem. Soc., Washington, DC, pp. 354–367. Roff, M. W. (1994). A novel lighting system for the measurement of dermal exposure using a fluorescent dye and an image processor. Ann. Occup. Hyg. 38, 903–919. Ross, J. H., Driver, J. H., Cochran, R. C., Thongsinthusak, T., and Krieger, R. I. (2001). Could pesticide toxicology studies be more relevant to occupational risk assessment?. J. Occup. Hyg. 45(suppl 1), 5–17. Ross, J., Chester, G., Driver, J., Lunchick, C., Holden, L., Rosenheck, L., and Barnekow, D. (2008). Comparative evaluation of absorbed dose estimates derived from passive dosimetry measurements to those derived from biological monitoring: Validation of exposure monitoring methodologies. J. Expo. Sci. and Environ. Epidemiol. 18, 211–230 March. Swan, A. A. B. (1969). Exposure of spray workers to paraquat. Br. J. Ind. Med. 26, 322–329. U.S. Environmental Protection Agency (1987). “Pesticide Assessment Guidelines, Subdivision U, Applicator Exposure Monitoring,” U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (1996). “Occupational and Residential Exposure Test Guidelines, OPPTS 875.1000, EPA 712C-96–261,” U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (1999). “Exposure Factors Handbook, EPA/600/C-99/001, February,” U.S. EPA, Office of Research and Development, Washington, DC.

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U.S. Environmental Protection Agency. (2007). “Review of worker exposure assessment methods,” FIFRA Scientific Advisory Panel Meeting, held at the EPA Conference Center, Arlington, Virginia. 9–12th January 2007. Van Dyk, L. P. and Visweswariah, K. (1975). Pesticides in air: sampling methods. Residue Rev. 55, 91–134. Van Heemstra-Lequin, E. A. H., and Van Sittert, N. J. (eds.) (1986). Biological monitoring of workers manufacturing, formulating and applying pesticides. Toxicol. Lett. 33, 1–236. Van Welie, R. T., van Duyn, P., Brouwer, D. H., van Hemmen, J. J., Brouwer, E. J., and Vermuelen, N. P. (1991). Inhalation exposure to 1,3dichloropropene in the Dutch flower-bulb culture. Part II. Biological monitoring by measurement of urinary excretion of two mercapturic acid metabolites. Arch. Environ. Contam. Toxicol. 20, 6–12. Versar. (1991). Pesticide Handlers Exposure Database (PHED). Data Entry Diskette User’s Guide, version 1.0. Report prepared by Versar Inc., Springfield, VA for the U.S. EPA, Office of Pesticide Programs, Health Effects Division, Occupational and Residential Exposure Branch.

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Vincent, J. H. and Mark, D. (1987). Comparison of criteria for defining inspirable aerosol and the development of appropriate samplers. Am. Ind. Hyg. Assoc. J. 48, 451–457. Wang, R. G. M., Franklin, C. A., Honeycutt, R. C., and Reinert, J. C. (1989). “Biological Monitoring for Pesticide Exposure: Measurement, Estimation and Risk Reduction,” ACS Symposium Series, Vol. 382. Am. Chem. Soc., Washington, DC. Wolfe, H. R. (1976). Field exposure to airborne pesticides. In “Air Pollution from Pesticides and Agricultural Processes” (R. E. Lee Jr., ed.). CRC Press, Cleveland, Ohio. Woollen, B. H. (1993). Biological monitoring for pesticide absorption. Ann. Occup. Hyg. 37, 525–540. World Health Organization. (1975). Survey of Exposure to Organophosphorus Pesticides in Agriculture. Standard Protocol, VBC/75.9. WHO, Geneva. World Health Organization. (1982). Field Surveys of Exposure to Pesticides. Standard Protocol, VBC/82.1. WHO, Geneva.

Chapter 53

Operator and Field Worker Occupational Exposure Databases and Modeling Curt Lunchick1, Jeff Evans2, Seshadri Iyengar3, Frank Selman4 and Heinrich Wicke3 1

Bayer CropScience, Research Triangle Park, North Carolina U.S. Environmental Protection Agency, Washington, DC 3 Bayer CropScience, Monheim, Germany 4 Dow AgroSciences, Indianapolis, Indiana 2

53.1 INTRODUCTION The use of pesticides to control insect pests, fungal diseases, and undesired plant pests (i.e., weeds) is an integral part of modern agricultural practices. Because pesticides are biologically active agents, it is essential to determine the human health risks incurred by the workers involved in the use of pesticide products. To quantify the risks associated with pesticide use, it is important to understand not only the toxicity of the particular active ingredient, but also the actual magnitude of the anticipated exposure. In this chapter an overview of the databases currently being used to make regulatory decisions and the transition to new models both in North America and Europe is presented. A discussion on the models and databases developed to predict exposure to re-entry workers is also covered. In this chapter a database is defined as a collection of worker exposure data from exposure monitoring studies that are contained in a program that permits the user to evaluate the data in the database. Databases should contain exposure data from a large set of field studies that form the appropriate basis for predictive extrapolation. A model is defined as a program containing exposure algorithms used to predict exposure for defined use scenarios. While a model does not contain actual data, as a database does, the predictive algorithms were developed from extensive evaluation of existing exposure data. Human exposures to pesticides may occur during worker contact involving principally the dermal and inhalation exposure routes. Worker populations that are routinely exposed to pesticides include agricultural handlers (or “operators” in Europe) involved in treatment of field crops, greenhouse crops, vineyards, and orchards; professional grounds applicators (e.g., parks and roadsides); lawn care professionals; structural and commercial Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

applicators (e.g., for factories, food processing plants, hotels, hospitals, other institutions, offices, and residences); and field workers (e.g., during harvesting or canopy management of treated crops) (Driver and Whitmyre, 1997; Maddy et al., 1990; U.S. EPA, 1984). In many cases there are two distinct operations in the application of pesticide products. These operations are the mixing/loading that involves handling the concentrated product and diluting it in preparation for application, and the actual application of the diluted spray solution to the intended target (van Hemmen, 1992). The requirement to quantify worker exposures to active ingredients during use of pesticide formulations is an integral part of risk assessments associated with the pesticide registration process in the United States, Canada, the European Union (EU), and other countries (e.g., Australia, Brazil). The risk assessment process enables regulatory agencies and the agrochemical industry to predict the extent of risk of adverse human health effects associated with the use of a given pesticide under specific use conditions. Because the evaluation of risk requires knowledge of both exposure and toxicity, exposures to the active ingredient associated with a given pesticide formulation must be assessed. This chapter provides a summary of the most commonly used worker exposure databases and models for mixing/loading and application of pesticides. The development of a new generic database for field worker exposures during re-entry activities (e.g., for scouting or harvesting) following application is also presented.

53.2 BASIS FOR GENERIC MODELING The basic premise of the generic modeling approach is that worker exposures are a function more of the work activity, 1139

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application equipment, formulation type, packaging type, level of clothing, total amount of active ingredient handled, and individual work practices than of the specific physicalchemical properties of the active ingredient. An analysis of the data from relevant literature studies indicated that where data were available for a given use pattern (e.g., mixing/loading and groundboom application of liquid formulations), the magnitude of exposure was independent of the physicochemical nature of the active ingredient (Day, 1991). Thus, measured dermal and inhalation exposures from a given set of studies on surrogate active ingredients can be used to approximate the worker exposures to a given active ingredient under similar use conditions. A number of field studies of varying quality have been conducted in the United States, Canada, Europe, and elsewhere that provide surrogate data for estimation of worker exposures to pesticides. Since 2002 the Agricultural Handlers Exposure Task Force (AHETF) has been conducting worker exposure studies using a standard criterion to minimize inconsistencies in study designs between studies and facilitate the comparison of exposure data between studies. Occupational exposure data for selected pesticides can be found in scientific publications, registration standards, and special review documents published by the U.S. Environmental Protection Agency (EPA) and evaluations by California’s Department of Pesticide Regulation, Worker Health and Safety Branch (e.g., Curry et al., 1995; Ecobichon, 1999; Fong and Krieger, 1988; Franklin and Worgan, 2005; Honeycutt et al., 1985; Honeycutt and Day, 2001; Krieger et al., 1990; Mehler et al., 1991; Nutley and Cocker, 1993; Plimmer, 1982; Rech et al., 1988; Saleh et al., 1994; Thongsinthusak et al., 1993; U.S. EPA, 1997; van Hemmen, 1992; Wang et al., 1989). Although formulation-specific agricultural worker exposure studies (e.g., mixer/loader, applicator, or harvester) may sometimes be necessary for pesticide registrations and reregistrations, exposure data on surrogate compounds for a given worker exposure scenario are often accepted. Selected exposure monitoring data form the basis for databases and models that allow estimation of exposures for specific application methods, use conditions, and formulation types. A generic exposure database provides several advantages to registrants, in that (1) they can use the database as a tool to estimate potential exposures and associated risks early in the product development stage and (2) they can use the surrogate data in the database to support registration and reregistration submissions to regulatory agencies, thus preventing expenditure of resources for generating compound-specific exposure data and reducing the time to registration or reregistration approval. Because of the larger pool of data that a generic database provides, greater reliability of the exposure estimates results and a better understanding of variability in the exposure estimates are obtained. In addition, a generic

Hayes’ Handbook of Pesticide Toxicology

database allows examination of the physical parameters and use conditions/work practices that affect exposure. Furthermore, a generic database provides a common basis for both industry and the regulatory agencies to arrive at the same exposure estimates using a common data source and method. In the European and North American generic databases and models, the exposure data are provided in generic form and are incorporated into the exposure model or summarized in normalized units [e.g., g/lb or kg of active ingredient (a.i.)]. The North American regulatory agencies (U.S. EPA, California EPA, and Health Canada) and European regulators have long endorsed this policy of using “surrogate” or “generic” exposure data for estimating worker exposures for the same work activity and work conditions, because it is impractical to conduct a field exposure monitoring study for every pesticide formulation. Indeed, the current policy in the EPA is to rely on the surrogate worker exposure data in the Pesticide Handlers Exposure Database (PHED) unless there is something unique about the use pattern or formulation under evaluation or the data in PHED are considered inadequate for the use scenario being evaluated (Jeff Evans, personal communication). European models such as the Predictive Operator Exposure Model (POEM; Martin, 1986) and the German model (FBRC, 1993) have been available for several years; these models are based on broad generic default values. The Europeans recognized the limitations of these models for applicability to EU-wide conditions and have focused their efforts on developing a scenario-specific mixer/ loader/applicator exposure database. These concerns and resulting refinements are reflected in the more recent tool known as EUROPOEM (AIR, 1996). The Agricultural Re-entry Task Force (ARTF), an industry task force in the United States, has developed an exposure database for different types of re-entry activities. Each type of activity (e.g., weeding, thinning, reaching, and harvesting) and crop type combination are associated with diff­erent rates of contact with treated foliage, known as the transfer coefficient (cm2/h). The ARTF re-entry exposure database and the clustering of activities contained in the database were presented to the U.S. EPA Science Advisory Panel in December 2008, and transitioning in North America to the use of the ARTF re-entry database began in 2009 and the U.S. EPA now relies upon the ARTF data for calculating reentry worker exposure. Similarly, the Outdoor Residential Exposure Task Force (ORETF) has performed a number of studies to characterize exposures to lawn chemicals on treated turf; these studies are being used by the member companies and by the regulatory agencies in North America. The number of studies conducted by ORETF is significantly less than those conducted by ARTF or being conducted by AHETF and therefore the ORETF data are being used as individual surrogate studies rather than in a database.

Chapter | 53  Operator and Field Worker Occupational Exposure Databases and Modeling

53.3  The tiered approach As with other applications of the risk assessment process, a tiered approach to estimating exposures is often taken (Figure 53.1). Although there is no universal tiered approach to risk assessment of worker exposures, one commonly used system can be described as follows. The first tier (tier I) of the risk assessment process for worker exposure to pesticides can involve the use of generic exposure data from databases and conservative assumptions for unknowns (e.g., for dermal absorption). The databases used could include the PHED, developed and used in North America (U.S. EPA, 1995a), POEM, developed in the United Kingdom (Martin, 1986), or the German model (Lundehn et al., 1992). PHED provides actual measured dermal and inhalation exposure data that can be retrieved for specific subsetting conditions (e.g., open mixing or open cab groundboom application). POEM provides an estimate of exposure modeled from actual exposure monitoring data based on the type of formulation, application method, application rate, etc. All of these databases and models are founded on similar principles in that they rely on measured exposure data from various studies that are combined into data sets for each worker exposure scenario. The estimated exposure from the appropriate surrogate database/ model is then used in combination with the key toxicological benchmarks [e.g., no observed adverse effect levels (NOAELs)] to determine whether the exposure is below the AOEL (acceptable operator exposure level) or whether

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an ­ acceptable margin of exposure (MOE) exists for the worker exposure scenario of interest. If exposure data of adequate quality and quantity exist and if using upperbound assumptions such as area treated, application rate, or hours worked, the tier I assessment demonstrates acceptable risk; the risk assessment process can be completed. In North America, the U.S. EPA and Health Canada have developed surrogate user guides for PHED for use in tier I assessments (PMRA, 2002; U.S. EPA, 1998). These user guides provide default unit exposure estimates from PHED for different mixer/loader and applicator use scenarios. Therefore, at the tier I level, the PHED model does not need to be utilized to create scenario-specific exposure estimates, as the regulatory agencies have already provided default scenario specific exposure estimates. If an unacceptable risk is obtained with a tier I assessment, it may be necessary to proceed with a more refined exposure assessment or to collect formulation-specific exposure data (tier II). In the latter case, tier II typically involves collection of external dosimetry data for workers on the formulation of interest, for specific use patterns of interest, thus circumventing the need to use surrogate monitoring data (e.g., from PHED or POEM) because compound-specific exposure data are collected. There needs to be a justification as to why compound-specific exposure data are being collected. Justifications could include a determination that the surrogate data were insufficient to address the use pattern, the equipment used in the surrogate data set is sufficiently different from current application equipment, or the formulation is appreciably different from

Figure 53.1  Tiered approach to estimating exposure. AOEL, acceptable operator exposure level; MOE, margin of exposure.

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those in the surrogate data. Without an appropriate justification there would be no reason to conclude that additional exposure data would be different from the surrogate exposure data if one believes that exposure data are generic. The exposure monitoring studies have been conducted on a variety of pesticides using commercial applicators, farmers, field workers, pest control operators, professional lawn care operators, and homeowners engaged in normal work activities. Several general approaches to quantifying exposures to pesticides for these individuals have been used, including the use of (1) patch (e.g., gauze pad) dosimetry, (2) glove dosimeter/hand rinse techniques, and (3) whole body dosimetry. Hand washes, patch dosimetry, and whole body dosimeters (both outer and inner dosimeters) are methods for quantifying the amount of pesticide that contacts the skin or clothing of a worker and thus, when adjusted for protection of specific body parts by clothing or measured directly via inner dosimeters, provide a measure of external (dermal) exposure. Clothing protection factors are calculated by comparing inner dosimeter residues to total external residues (inner plus outer dosimeter residues) for each body part section. The use of whole body dosimeters, which are usually sectioned into standard body part areas (e.g., upper legs or lower legs) before extraction and analysis, eliminates the need for extrapolation from a small patch size to the whole body part. In addition, fixed location or personal air-monitoring devices have been used to characterize exposures via the inhalation route by collecting a known volume of air in the breathing zone of the worker and analyzing for the mass of pesticide of interest present. Because dermal exposures represented by the dosimetry data are external exposures, it is necessary to apply assumptions on clothing protection factors (if inner dosimeters are not utilized) and dermal absorption to estimate the absorbed dermal dose, unless a dermal toxicity study is available that provides the NOAEL for an applied dose. If a dermal toxicity study is not available, an additional tier II refinement might include producing a dermal absorption study, since the default in the absence of data is frequently 100%. If the tier II assessment still does not indicate acceptable risk, the absorbed dose can be measured directly by means of a biomonitoring study (tier III). Biomonitoring, also known as biological monitoring, typically uses the amount of pesticide (or its metabolites) detected in the urine of exposed individuals to obtain an accurate measure of the total amount of pesticide actually absorbed by the worker via all routes (inhalation, dermal, and incidental oral ingestion). Biomonitoring provides total absorbed dose (i.e., pesticide level in the body) as a result of the exposure at the body’s boundaries (e.g., skin or lungs), but it does not explain or separate the contributions of each specific exposure pathway. However, biomonitoring data negate the need for extrapolation from external dosimetry to internal

Hayes’ Handbook of Pesticide Toxicology

dose; thus, biomonitoring data provide a less conservative and more meaningful (direct) measure of exposure.

53.4  The pesticide handlers exposure database (PHED) 53.4.1  Historical Background In 1983, the Public Health and Toxicology Committee of the National Agricultural Chemicals Association (NACA) (now CropLife America) held a workshop on pesticide worker exposure. Following this workshop, the Subcommittee on Field Exposure Assessment was formed for the dedicated purpose of developing a standard protocol for conducting field studies of mixer/loader/applicator exposures (Day, 1991). At the first meeting of the subcommittee, Drs. Hackathorn and Eberhart from Mobay (now Bayer CropScience) presented findings from a literature search of relevant exposure studies. An analysis of the data from these studies indicated that where data were available for a given use pattern (e.g., mixing/loading and groundboom application of liquid formulations), the magnitude of exposure was independent of the physicochemical nature of the active ingredient (Day, 1991). Thus, they conjectured that if there were enough data available for a given use pattern, it should be possible to use these data as surrogate data that would be universally applicable for the estimation of exposures for that use pattern (Day, 1991). Therefore, new field studies would not need to be conducted on every formulation that was presented to the U.S. EPA for registration. The use of generic exposure data and the development of a worker exposure database were proposed and discussed at the 187th meeting of the American Chemical Society in April 1984. Three important papers were presented (Hackathorn and Eberhart, 1985; Honeycutt, 1985; Reinert and Severn, 1985), which proposed how a generic exposure database could be used in the risk assessment process (Lunchick et al., 1994). As a result of these proposals to develop a generic database for applicator exposures, a task force consisting of representatives from Health and Welfare Canada (now Health Canada), the U.S. EPA, and the National Agricultural Chemical Association (now CropLife America) was formed to oversee the design and development of what would ultimately become PHED. Industry agreed to allow proprietary exposure data to be included in the database as long as the identity of the specific chemicals was kept confidential. Because the exposure studies would represent a variety of study designs, there was a need to anticipate the various data set types and the data fields that would be necessary in PHED to hold the data. Furthermore, there was discussion as early as 1985 regarding the need to rank the data based on adequacy of quality assurance procedures, so that users would

Chapter | 53  Operator and Field Worker Occupational Exposure Databases and Modeling

be able to create subsets from the data based on quality assurance ranking or grade. A joint EPA–industry task force was formed, headed by the U.S. EPA. A generic database concept paper was presented at the Sixth International Congress of Pesticide Chemistry sponsored by IUPAC in Ottawa, Canada, August 10–15, 1986. Requests were made to industry members (through CropLife America) in 1987 for submission of studies to the task force (through the U.S. EPA) for entry into the database (NACA, 1987). Registrants were asked to sign a data compensation waiver and to enter key data from their studies on a form that would later be used by a U.S. EPA contractor as a data entry form. A number of years later, a data entry diskette was developed by a U.S. EPA contractor (Versar, 1991), which provided a mechanism for registrants to submit their data to the task force in electronic form.

53.4.2  Overview of PHED PHED was developed as a joint effort between the U.S. EPA, Health Canada, and CropLife America to provide dermal and inhalation data for assessing mixer/loader and applicator exposures during application of pesticides. This database was first released for general use as version 1.0 in May, 1992 (Lunchick et al., 1994). Since that time PHED has been used by registrants and government agencies to supplement or replace worker exposure studies, as a validation tool for field exposure data, for estimating exposures during the early stages of product development (e.g., registerability assessments for alternative types of formulations and levels of active ingredient), and for fulfilling data requirements imposed by regulatory agencies (U.S. EPA, Health Canada, and the California EPA). PHED, version 1.1 (U.S. EPA, 1995a), which was released in 1995, contains more data than version 1.0 (over 1700 records) on measured dermal and inhalation exposures and on various parameters that may affect the magnitude of exposures. Each data record represents one replicate for one worker involved in one work day or less of a given activity. There are four separate data files in PHED where the worker exposure data reside, depending on which work activities were monitored. These include the mixer/loader (MLOD file, 556 replicates), applicator (APPL file, 715 replicates), mixer/loader/applicator (MLAP file, 349 replicates), and flagger (FLAG file, 92 replicates) files. PHED can be used to develop unit exposures, typically g/lb or kg active ingredient (a.i.) handled, for a specific worker exposure scenario, and also permits the user to express the exposure normalized by hours or as raw exposure not normalized by any variable. PHED provides a useful tool for modeling and predicting potential pesticide exposures based on consideration of numerous factors, such as application rate, formulation type and packaging, mixing/loading methods, application methods and equipment, and the type

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of clothing or protective equipment used. Guidelines have been developed for proper use and reporting of PHEDderived data (U.S. EPA, 1995b).

53.4.3  PHED Grading Criteria Each data replicate contained within PHED is classified according to specific quality criteria. The data-grading scheme developed by the U.S. EPA, Health Canada, and Croplife America assigned letter grades (A through E) based on analytical quality. By basing the grading exclusively on analytical quality, all valid data points, regardless of the completeness of the study or the quality of its parts, are available for use. The guidelines for using PHED (U.S. EPA, 1995b) specify that grades A and B data should be used. Class A data are those data for which (1) the associated laboratory recovery is 90–110% (with a coefficient of variation of less than or equal to 15%), (2) the associated field recovery is 70–120%, and (3) the field recovery samples have experienced the same environmental conditions for the same duration as the field monitoring samples.1 Usually the worker exposure monitoring data will have been corrected based on field recovery unless the field recovery value was 90% or more. Class B data are those data for which (1) the associated laboratory recovery is 80–110% (coefficient of variation of 25% or less) and (2) the field recovery data are present and in the range of 50–120%. Class C data are those data for which (1) the associated laboratory recoveries are 70–120% (coefficient of variation of 33% or less) and (2) either the associated field recovery data are present and in the range of 30–120% or field recovery data are absent and the storage stability data are in the range of 50–120%. Class D data are those data for which (1) laboratory recoveries are 60–120% (coefficient of vacation of 33% or less) and (2) the field recovery or storage stability data are either present or missing. Data not meeting the criteria for classes A through D (e.g., if laboratory recovery data were not reported) are assigned to class E.

53.4.4  Guidance for the use of PHED The U.S. EPA first published guidance for the proper and appropriate use of PHED in 1995 with the release of PHED version 1.1 (U.S. EPA, 1995b). It is first of all necessary for exposure assessors to accurately define the exposure scenario based on the pesticide product label. Second, it is important for the exposure assessor to subset

1 

Field recovery samples that are spiked with active ingredient and then placed immediately in a cooler are not field recovery samples; rather, they measure storage stability.

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the data in a consistent and logical way. To achieve this end, the U.S. EPA guidance document (U.S. EPA, 1995b) indicated the key subsetting parameters for each of the four different data files. For example, mixer/loader data would typically be subset based on formulation type (e.g., wettable powder), packaging type (e.g., water-soluble packets), mixing procedures (e.g., open versus closed), and in some cases based on a specific range of total amount of active ingredient mixed. Applicator data would typically be subset by formulation type (e.g., formulations applied as a solid or formulations applied as a spray), application method, and cab type (e.g., “open” versus “closed”). In essence, the selected subset should mimic as much as possible the actual use scenario intended for analysis. The acceptability of a data subset from PHED should be determined by six factors: sample size, data quality, duration of sampling period, key body regions, clothing scenarios, and likelihood of incidental contact. EPA Series 875 Group A Applicator Exposure Monitoring Guidelines specify that a minimum of 15 replicates be present in an exposure data set; however, the minimum acceptable number is determined on a case-by-case basis. The data set size of 15 was initially intended as guidance for the conduct of a single study and was not intended to represent the target sample size of a subset from a generic database such as PHED. Sample size requirements are affected by the purpose of the assessment, the quality of the data, and the anticipated variability in exposure data for the worker exposure scenario. The quality of the data, which is assigned separately to airborne, hand, and other dermal exposures for each worker replicate, would normally be limited to data quality grades A and B (see preceding discussion), although exceptions may be made for some use scenarios in which the data subset is expanded to include C grade data in order to obtain a data set of adequate size. The sampling period for each worker replicate corresponds to the duration of a given work activity (e.g., mixing/loading or application). Exposure data should be collected over a period of time that is representative of typical work practices. For some work activities, typical durations may be short (e.g., 20 min for mixing/loading) or represent most of the work day (e.g., 4–6 h for application). Data sets from PHED involving longer durations of time are preferred to those associated with short sampling periods, because they are more like a normal work day and tend to have fewer nondetects. Key body regions relevant to the work activity must be represented in the data set. For example, an exposure data set for mixing/loading that contains none or just a few replicates of data for hand exposure would not be considered a valid data set because hand exposures typically make up the majority of total exposures for mixing/ loading activities. Either the clothing scenario for which the data set is obtained from PHED as reflected in the dosimeter location (e.g., outside of clothing, under normal clothing, or under normal and protective clothing) should

Hayes’ Handbook of Pesticide Toxicology

be ­consistent with pesticide label requirements, or the data (e.g., for total deposition on the skin) should be adjusted to reflect the anticipated clothing scenario using standard clothing protection factors. One or a few replicates of data that produce a body part-specific exposure that is dramatically higher than the central tendency value for that body part may indicate that incidental contact has occurred (e.g., a spill on clothing). It requires careful judgment and justification to determine whether specific worker replicates be deleted from the data subset before analysis. As PHED version 1.1 was used to support exposure assessments, it became evident to the assessors within the EPA that the subset definitions for specific use scenarios that were being submitted by registrants or created by the EPA assessors were essentially the same. When PHEDbased exposure assessments were submitted by pesticide registrants the EPA had to verify the accuracy of the submitted assessment, a time-consuming activity. Furthermore, the differences in the unit exposures between different PHED subset criteria for a given scenario were generally not large, because the subset criteria were usually similar. For example, open-pour liquid mixer/loader assessments typically specified liquid formulations, open-pour mixing procedures, and data grades A and B for dermal, hands, and inhalation data. Differences in the subset definition included variables such as spray tank size, amount of spray mixed, or amount of active ingredient handled. In August 1998 the EPA issued the PHED Surrogate Exposure Guide (U.S. EPA, 1998). The guide provided unit exposure estimates expressed in mg/lb a.i. for three clothing scenarios (no clothing or potential dermal exposure, single layer of clothing without protective gloves, and a single layer of clothing with protective gloves) and in g/lb a.i. for inhalation exposure. The EPA’s confidence in the data for each exposure estimate was also provided. Confidence was rated as high, medium, or low based on the number of records in the data set and the data quality grade. High-confidence data sets contained at least 15 records of grade A or B data for each PHED body area. Medium-confidence data sets contained at least 15 records for each PHED body area but also included C grade data to obtain the 15 or more records. The low-confidence data sets contained either less than 15 records for at least one body area or contained D or E grade data. It was the EPA’s anticipation that additional exposure data were to be submitted for medium- and low-confidence data sets so that they would eventually be upgraded to high-confidence data sets. The surrogate guide also established 37 different use scenarios. The use scenarios are provided in Table 53.1. Health Canada released the Canadian PHED Tables (PMRA, 2002) in 2002. The Canadian PHED Tables provided tier I PHED unit exposure estimates for use in exposure assessments submitted for Canadian regulatory purposes. The Canadian PHED Tables are similar to the EPA’s, but the Canadian tables provide more clothing options, contain fewer use scenarios, and all units are metric.

Chapter | 53  Operator and Field Worker Occupational Exposure Databases and Modeling

Table 53.1  PHED Surrogate Exposure Guide Scenarios Scenario no.

Scenario description

EPA confidence level

1

Dry flowable: open mixing

High

2

Granular: open loading

Medium

3

All liquid formulations: open mixing

High

4

Wettable powders: open mixing

Medium

5

Wettable powders: water-soluble bags

Low

6

All liquid formulations: closed systems

High

7

Aerial fixed wing: liquid sprays

Medium

8

Aerial fixed wing: granules

Low

9

Rotary application

Extremely low

10

Aerosol application (spray can)

High

11

Airblast application: open cab

High

12

Airblast application: enclosed cab

High

13

Groundboom application: open cab

High

14

Groundboom application: enclosed cab

Medium

15

Solid broadcast spreader: open cab

Low

16

Solid broadcast spreader: enclosed cab

High

17

Hand-dispersed granular bait

Medium

18

Low-pressure handwand application

Low

19

High-pressure handwand application

Low

20

Backpack application

Low

21

Lawn handgun sprayer

Low

22

Paintbrush application

Low

23

Airless sprayer

High

24

Rights-of-way sprayer

Low

25

Flagger during liquid application

High

26

Flagger during granular application

Potential dermal exposure data only

27

Open loading/open cab airblast

Potential dermal exposure only

28

Open loading/open cab groundboom

Medium

29

Open loading/enclosed cab groundboom

Medium

30

Granular loading/belly grinder application

Medium

31

Open loading/push type spreader

Low

32

Open pour liquid mixing/low-pressure hand spray

Low

33

Open pour wettable powder mixing/low-pressure hand spray

Medium

34

Liquid open pour mixing/backpack application

Low

35

Liquid open pour mixing/high-pressure hand spray

Low

36

Liquid open pour mixing/hose end sprayer

Low

37

Liquid open pour mixing/termiticide injection

High

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53.4.5  Limitations of PHED Because PHED was developed in the late 1980s and the early 1990s, some of the studies in PHED were not conducted according to good laboratory practices (GLPs), and some of the more recent studies, conducted under GLPs, are not in the database. PHED is populated with studies containing a high percentage of nondetects, with minimum detection limits for many studies that range over several orders of magnitude; thus, for some subsets, the exposure estimates are mainly driven by the limits of quantification. Because of the wide differences in study designs and detection limits, extremely high variability occurs in the aggregated data in the exposure subsets from multiple studies. Thus, exposure statistics other than the central tendency may have limited meaning (van Hemmen, 1992). Many replicates are of such short duration or involve handling of such small amounts of material that they are not generally applicable to exposures associated with more typical worker task durations or more typical amounts of active ingredient handled per day, respectively. Because chemical-specific properties (e.g., vapor pressure) may affect penetration of worker clothing, the protection provided by personal protective equipment (PPE) or engineering controls may be underestimated by the use of PHED data. These limitations have been recognized by both the North American regulatory agencies and pesticide registrants. It was the recognition of these limitations that led to the formation of AHETF and the ongoing development of a replacement agricultural handler exposure database.

53.5  European union models and databases 53.5.1  Overview of the German and UK Operator Exposure Models (BBA Model and UK-POEM) The German BBA model and the UK-POEM (Predictive Operator Exposure Model) are the two operator exposure models officially recommended by Council Directive 91/414 for lower-tiered risk assessment in the EU regulation.

of agrochemicals in Germany but also has been adopted at the EU level as one of the two models recommended for inclusion of compounds in Annex I. The model provides single study descriptions and individual results, thus allowing a detailed understanding of how the model defaults are derived. The BBA model is embedded in a comprehensive concept of operator protection. The principle of this concept consists of two key elements: specific instructions of operator protection derived in a first step from toxicological classification and in a second step from risk assessment. Specific instructions consist of (a) elements of protective gear, (b) type of product (undiluted, diluted), and (c) use conditions (use pattern and application technique). For risk assessment, operator exposure is estimated with the exposure algorithms provided in the BBA model and then compared with the tolerable exposure. Excedance of the tolerable exposure requires implementing model instructions to reduce exposure via additional PPE using provided reduction coefficients in the revised calculation. The model takes into account the three major routes of exposure: dermal (D), oral (O), and inhalation (I). The oral exposure has experimentally been accounted for via inhalation exposure and is therefore not addressed separately. Further differentiation is made for the purposes of analyzing the dermal exposure. Therefore, the operator body is divided into three parts: hands (H), head (C), and rest of body (B). The model differentiates exposure during handling the concentrate (M: mixing/loading) and exposure during handling the spray dilution (A: application). During mixing/loading, exposure to the hands contributes the dominant share to the total dermal exposure. Therefore, only inhalation and hand exposure are considered in this working step. According to the different working steps, the total exposure (E) consists of the following partial exposures:

The German BBA model was compiled in the late 1980s by the BBA (Biologische Bundesanstalt für Land- und Forstwirtschaft), the BGA (Bundesgesundheitsamt), and the German Agrochemical Manufacturers Association. The database of the model consists of (non-GLP) data from experimental pesticide studies conducted by the German pesticide manufacturing industry. The effort resulted in a generic exposure model (Lundehn et al., 1992) that not only is used in the regulatory process for the registration

(1)



Each partial exposure is calculated with experimentally determined specific exposures (D* and I*), the daily treated area, and the dose rate:

53.5.1.1  German BBA Model

E  D M(H)  I M  D A (HCB)  I A

D M(H)  D*M(H)  A (area )  R (dose rate), etc.



(2)

Specific exposure values are normalized default values based on units of weight per amount of active substance handled (mg/kg a.s.).2 They are the geometric mean values of the respective data range of the partial exposures and are available for a range of partial exposures that are composed of exposure route (dermal or inhalation), body 2 

Note that the term active substance (a.s.) is commonly used in Europe and is interchangeable with the term active ingredient (a.i.) used in North America.

Chapter | 53  Operator and Field Worker Occupational Exposure Databases and Modeling

part (hands, head, or body), working step (mixing/loading or application), the type of formulation (liquid or solid, the latter either as WG or WP), the application technique (tractor sprayers or handheld sprayers) and the cropping pattern (arable crop or high crop). The daily treated area considered for tractor-mounted/trailed boom sprayers is 20 ha/day and for tractor-trailed air-assisted sprayers is 8 ha/day. Handheld sprayers are considered with 1 ha/day in the calculation. The maximum label rate is used for each. The exposure is modeled following consecutive steps. In the first step, exposure of an unprotected operator is calculated supposing a moderately dressed person with half of the upper arms, forearms, thighs, and lower legs unprotected. If the estimated exposure exceeds the tolerable exposure, the exposure of a protected operator is calculated in a second step. The mitigation of exposure includes wearing single additional PPE or a combination of various PPE such as standard protective coverall, universal protective gloves, chemical protective clothing (type 3), head gear, hood and visor, and/or inhalation protection via particle filtering half mask or combination mask against gases, vapors, and particles. Exposure is then calculated using individual reduction coefficients for each PPE as a measure of protection both for the dermal and for the inhalation route. For risk assessment, the final exposure is compared to a tolerable dose either in a route-specific or in a route-to-route approach. The route-specific approach compares each exposure route separately with the tolerable doses (Dtol and Itol), which are based on a lowest NOAEL determined for the route. The sum of the route quotients results in the total degree of exposure relative to acceptable exposure: E

D D

tol



I I

tol

 O    tol   O 

(3)

A total degree of exposure of 1 is acceptable. The predominantly used method, however, is the routeto-route approach, as most existing toxicity databases contain studies involving predominantly the oral route. Dermal (DE) and inhalation exposure (IE) is therefore transformed into absorbed doses using percutaneous (PA) and inhalation absorption rates (IA). An adult body weight of 70 kg is assumed. The sum of the absorbed rates, the total systemic exposure (SE), is then compared with the AOEL.

SE  (DE  PA)  (IE  IA)

(4)

A systemic exposure of 100% of the AOEL is acceptable.

53.5.1.2  UK-POEM In May 1985, the Joint Medical Panel of the Scientific Subcommittee on Pesticides and the British Agrochemical Association Toxicology Committee decided to review the available pesticide worker exposure data to determine

1147

the extent to which they would support a generic exposure model (Martin, 1986). The result of the effort that foll­owed this decision to review the available exposure data is the UKPredictive Operator Exposure Model (POEM). The model is compiled from experimental pesticide and tracer studies performed in the 1970s/1980s in the United Kingdom under UK conditions and was made available in the late 1980s to estimate the levels of exposure likely to be incurred by operators (mixer/loader/applicators) applying pesticides. Like the German BBA model, the UK-POEM was adopted by the European Commission for pesticide risk assessment at the European level although the exposure data may not be universally relevant to application conditions in every European country. A complete discussion of POEM has been provided by Martin (1986, 1990), and Hamey (1992) provides the user’s guide for the model. POEM is based on limited generic exposure monitoring data extended to cover a variety of formulation types and use scenarios. However, no study descriptions are given and only summary data are provided. Also, in the original version no data were available to calculate exposure during mixing and loading (M/L) of solid formulations. A new version of the UK-POEM was therefore compiled by Pesticides Safety Directorate (PSD) and published in 2003 with a set of changes and updates. For solid formulations: Inclusion of dermal and inhalation exposure data for mixing/loading of solids from the German BBA model l Differentiation of formulation types WP or SP, WG or SG l Inclusion of WB (water-soluble bags) l Additional PPE during M/L for respiratory protection (FFP2 and FFP3 mask) including new mitigation factors l New mitigation factor for gloves (1%) l Additional application method home garden sprayer (5-l tank) outdoor, low level. l

For liquid formulations: Additional 20-l container (with narrow and 63-mm closure) l Additional PPE: impermeable coverall for handheld applications l Additional application method home garden sprayer (5-l tank) outdoor, low level. l

Finally, further changes were made by PSD in 2008, and the latest version of the UK-POEM was published on the PSD website. The newest version takes into account new values for inhalation exposure during mixing/loading solid formulations for application using vehicle-mounted equipment: A new value for inhalation exposure when mixing/loading WG from EUROPOEM

l

Hayes’ Handbook of Pesticide Toxicology

1148

A new value for inhalation exposure when mixing/loading WP formulations from PHED.

l

Variables that describe the use conditions are utilized in POEM to predict daily exposure. These variables include formulation type and active substance concentration, main solvent, container size and design, method of application, application rate, spray volume, PPE recommendation, and the daily area treated. The application methods addressed in POEM include tractor-mounted/trailed boom sprayer either with hydraulic nozzles or rotary atomizer, tractor-mounted/trailed broadcast air-assisted sprayers (very low volume, low volume, and high volume), low-level handheld outdoor sprayers, low-level and high-level handheld rotary atomizers, and low-level home garden sprayers. The model contains algorithms and structured calculation procedures in spreadsheet format (Hamey, 1992). The impacts of PPE and formulation differences are also addressed. Exposure values are normalized in units of volume that impinge on the operator, for example, volume of product (ml) per number of operations used for the calculation of exposure during mixing/loading or volume of spray per spraying time (ml/h) used for the calculation of exposure during application. The measured exposure values in the underlying database are ranked in different classes, and the exposure values presented in the model are the upper end of the class that contains the 75th percentile value. These upper end values are high end percentiles typically in the 80th– 90th percentiles depending on the application scenario. Exposure values for mixing/loading solid formulations follow the normalization approach of the exposure model from with they are adopted (mainly from the BBA model, two inhalation exposure values from the EUROPOEM and PHED), which is units of weight per amount of active substance handled (mg/kg a.s.). These exposure values are 75th percentile surrogate values (or maximum values for small data sets (i.e., less than 10 measured data values). POEM has been approved by the Scientific Subcommittee and the Advisory Committee on Pesticides and is used for regulatory purposes by the Pesticides Safety Directorate in the United Kingdom. The estimation of operator exposures using POEM is divided into two parts: (1) estimation of exposures associated with handling the concentrated formulation (e.g., during mixing and loading procedures) and (2) estimation of exposures associated with actual application of the diluted formulation. For exposures during the mixing/loading operation, POEM assumes that only hands are contaminated; generic data in POEM used for mixing/loading liquid formulations are based on exposures during container pouring tests rather than on exposures during actual spray tank loading. These data indicate the dependence of operator exposure during mixing/loading on the size of the pesticide formulation container and its neck aperture width. Data used for

mixing/loading solid formulations are based on exposures during actual spray tank loading. For the application of pesticide formulations, both dermal and inhalation exposures are considered in POEM. A number of factors are considered for evaluating the total exposure. These include, for example, the work rate (i.e., area treated per day), the distribution of pesticide contamination on the worker (percentage of total contamination on hands, trunk, and legs), the anticipated penetration of clothing (assigned separately for hands, trunk, and legs), the duration of exposure to the spray (h), and the percutaneous absorption of the active ingredient. It is assumed that vehicle-mounted hydraulic-boom applications are made to 50 ha/day.3 For air-assisted applications to orchards, the area treated is assumed to be 15 ha/day. Handheld applications are associated with a maximum work rate of 1 ha/day. The duration of exposure to spray, corresponding to duration of the application work task, is typically assumed to be 6 h/day, at least for vehiclemounted applications and handheld applications. Potential inhalation exposures to formulations ranging from 0.005 to 0.05 ml/h are assigned in POEM depending on the type of application method, protective measures (e.g., use of cab), and application volume. POEM combines the dermal absorbed dose and the inhalation exposure to obtain a total absorbed dose [mg/kg body weight (bw)/day], based on a 60-kg adult body weight. The total absorbed dose is then compared to an AOEL.

53.5.1.3  Limitations of Both Models Data were obtained in the late 1970s to 1980s and may therefore not represent modern application technique; no modern studies are included. l None of the studies were performed according to GLP. l The databases contain nondetects, and some exposures are driven by the level of quantification (LOQ). l UK-POEM and the German model assume that only the hands are exposed during mixing and loading. l UK-POEM assumes no inhalation exposure during mixing and loading of liquids. l There are no exposure data in UK-POEM for handheld upward spraying, for example, high crops. l Three are no exposure data in the German model for handheld downward spraying, for example, for weed control. l There are no exposure data in UK-POEM for spraying of grapes. l UK-POEM tends to overestimate exposure. The Advisory Committee on Pesticides’ comparison to biological monitoring data has shown that the model estimates always l

3

 This is reasonable for grain crops, but a lower value for the area treated per day may be justified for other crops (Hamey, 1992).

Chapter | 53  Operator and Field Worker Occupational Exposure Databases and Modeling



l



l



l



l

exceeded the measured data. This observation is supported by findings of the Health and Safety Laboratory (HSE)4 that demonstrate significant overestimation by the generic figures given in the UK-POEM, especially for dermal exposure, independent of the spray volume. Various types of pesticide application are not included in both models, for example, dusting, fogging, and dipping. Exposure data are derived from exposure studies performed either in Germany or in the UK. Both models may therefore be used with reservation for crops in other countries with different climate or cropping pattern, for example, citrus or olives. There are no data for greenhouse application in both models. A new model is under preparation and a draft version is proposed by the European Crop Protection Association (ECPA). There are no data for seed treatment. This application is considered in a separate operator exposure model (SeedTropex).

53.6  The agricultural handlers exposure database (AHED) The AHETF, in cooperation with the European Crop Protection Association (ECPA) Occupational and Bystander Exposure Expert Group and the Antimicrobial Exposure Assessment Task Force (AEATF), has developed database software, accessible via a web-based server, for handling the data entry, calculations, and data analysis of handler exposure data being developed by the three task forces. Thus, a common database software tool will be used for risk assessments both in North America and in Europe. The data generated by each task force will be uploaded into each task force’s version of AHED to address data ownership. In this way the software is similar, but the data will only be that which is applicable to each task force. AHED was created in response to the recognition of the limitations of the existing North American and European handler exposure databases and models and the ongoing conduct of exposure studies to address the data gaps and deficiencies previously discussed. The new database was developed by infoscientific.com in Visual Studio.NET and written in Visual Basic and Visual C#. The database itself is MSDE 2000. The developers of AHED are committed to making the program itself freely available, but the data contained within the database are subject to ownership rights in the jurisdictions where it will be used. The database is organized to contain tables providing information on the worker, application procedures, mixing 4

 Technical Development Survey: Exposure to chlorpyrifos in orchard spraying; published by Health and Safety Executive, 9/98; Health and Safety Executive 1995 and 1996.

1149

procedures, product information (except for the product name and active ingredient), PPE worn, monitoring media and supporting analytical chemistry, and the exposure data for each individual monitoring unit in AHED. The user will be able to define scenario use criteria to create a subset of exposure data specific to the pesticide use of interest. The exposure data can be expressed as the actual exposure measured during the monitoring period, or the exposure data can be normalized based on the amount of active ingredient handled or by time. As with PHED, the user also has the capability to define the clothing worn for a specific use scenario. AHED can perform limited descriptive statistical analysis of the data and provides the arithmetic and geometric mean exposure estimates of the subset with the appropriate standard deviation. In addition, the database also provides 50th, 75th, 90th, and 95th percentiles of the exposure distribution plus the minimum and maximum values. The statistical capabilities were intentionally limited in AHED because the user can export the data from a subset into Excel and then use any available statistical program to conduct additional statistical analyses. AHED is also designed to export tables, graphs, and information from the user’s subset directly into a Microsoft Word document. Examples of AHED screens are presented in Figures 53.2 and 53.3. AHED version 1.0 was released to members of AHETF in 2007 for learning purposes and internal use. Based on use of version 1.0, additional programming changes were made for version 1.1. The programming of AHED version 1.1 is undergoing quality assurance auditing during the summer of 2009 with an intended release by the end of 2009. The populating of AHED with data from the respective task force data development efforts is an ongoing process that will continue after the release of version 1.1. AHETF is working in cooperation with the U.S. EPA, the California Department of Pesticide Regulation, and Health Canada on the development of additional exposure data and entry of data into the AHETF-AHED. The transition from PHED to the AHETF-AHED database is intended to occur on a scenario basis as AHETF and the North American regulatory agencies conclude that the data in a given scenario are complete. Similarly, ECPA will work with the European Food Safety Agency (EFSA) and key member states on a similar process of populating the EU-AHED with exposure data for use within the EU in assessing operator exposure under Directive 91/414.

53.7  Re-entry exposure database Unlike mixer/loader/applicator exposures, for which PHED has been an available tool for obtaining normalized exposure to pesticides for specific use scenarios (U.S. EPA, 1995a), a completed publicly available database on worker re-entry exposures did not exist until recently. The general

1150

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Figure 53.2  AHED data query page.

Figure 53.3  AHED unit exposure output.

concepts of such a database and how the information in it would be applied have been described by others (Honeycutt, 1985; Nigg et al., 1984; Zweig et al., 1985). The ARTF, whose member companies manufacture and/or distribute pesticides, has developed a worker re-entry database for use by its member companies.

The ARTF database contains generic transfer coefficients (TCs) that are representative of specific crop types and worker activities, based on actual field studies sponsored by the ARTF or purchased by the ARTF. Generic transfer coefficients are used to estimate worker re-entry or postapplication exposures. The ARTF re-entry database

Chapter | 53  Operator and Field Worker Occupational Exposure Databases and Modeling

contains data on worker dermal and inhalation exposures, dislodgeable foliar residues, site locations, meteorological conditions, and other ancillary information such as formulation type, method of application, and re-entry times and conditions. The database permits the user to obtain transfer coefficients on a whole-body and a body-part-specific basis by clusters. Each ARTF TC cluster consists of several crop/activity/height/foliage combinations that have a similar pattern and magnitude of exposure potential and that are represented by transfer coefficients measured within re-entry exposure studies on certain crop/activity/height/ foliage combinations within the cluster.

53.7.1  ARTF Cluster Survey To determine which crops and activities needed to be represented; ARTF conducted a survey of agricultural experts in 1995 to determine the activities performed within some 90 crops and what type of foliar contact was expected for the various crop/activity combinations (Bruce and Korpalski, 2008). The list of crops was derived from the 1992 U.S. Census of Agriculture and the 1990 Canadian Census of Agriculture. Experts that participated in the survey included crop specialists from universities, cooperative extension agents, crop consultants, farm managers, pest control advisors, entomologists, and pesticide suppliers throughout the United States and Canada. A grower survey was planned as a follow-up that would involve a much larger sample size and provide the basis for crop/activity clustering. The expert survey was conducted first so that the overall approach to collecting this information could be tested and the questionnaire could be refined based on expert input. The expert survey included focus group discussions as well as one-on-one phone interviews, and participants were asked to fill out a detailed questionnaire for the specific crop(s) with which they were familiar. The following information was collected through these crop-specific questionnaires and was intended for ARTF use in clustering re-entry activities: re-entry activities that are performed on the specific crop l whether contact with foliage occurs with each activity l crop height at the time of each re-entry activity, as one of three categories:  30 inches, 30–48 inches, or 48 inches l degree of foliage present at the time of each re-entry activity, as minimum, moderate, or full l relative frequency of contact with treated foliage for each of nine body parts during each re-entry activity, as a numeric index value from 1 to 3 representing no, occasional, or continuous contact l relative degree of contact with treated foliage for each of nine body parts during each activity, also as a l

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numeric index value from 1 to 3 representing no, partial, or complete contact l groups of crops that reflect similar re-entry activities and contact potential. Despite its subjectivity, the results of the expert survey supported collection of information on re-entry activities and foliar contact using a survey format. In addition, the clustering of the crop/region/activity/height/foliage (C/R/ A/H/F) combinations based on contact potential (CP) similarity used an established multivariate technique. Based on the expert survey experience, a very similar grower survey was designed to collect a larger sample of CP estimates (Thompson, 1998). Several improvements were made to the grower survey based on the expert survey experience, including the following: The crop list (C) was expanded to include six additional crops suggested by experts. l Regions (R) were redefined as the 13 growing regions established by the EPA (EPA OPPTS Residue Chemistry Guideline 860.1500: Crop Field Trials) plus three regions representing Canada. l A quota system was established to ensure an adequate sample size for each region/crop based on important regions identified by acreage data and high contact activities identified by the expert survey. l The activities (A) list was modified slightly (e.g., hand and mechanical weeding were separated; planting and transplanting were combined for some crops). l Crop height (H) definition was changed from an absolute scale to a relative scale of low, medium, and high, defined as one-third, two-thirds, and full height relative to crop maturity, respectively. l Degree of foliage (F) present at the time of each reentry activity was characterized as minimum, moderate, or full, as in the previous survey. l Several questions were slightly reworded for clarity. l

ARTF began its field testing program late in 1996, soon after grower survey data were available. Studies were conducted according to a standardized generic protocol for field reentry monitoring using passive dosimetry that was compliant with the relevant EPA guidelines (U.S. EPA, 1996) and approved by representatives from the EPA, Health Canada, and the California Department of Pesticide Regulation. The passive dosimetry exposure monitoring methodologies contained in those guidelines were reviewed thoroughly at a meeting of the U.S. EPA Scientific Advisory Panel in 2007 (FIFRA Scientific Advisory Panel, 2007). No significant changes to the passive dosimetry exposure monitoring techniques used by ARTF have been recommended. These techniques are still in use, largely unchanged from ARTF’s procedures. ARTF completed a comprehensive report on its ­transfer coefficient database in 2003 that contained a proposal

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for 25 clusters of C/A/H/F combinations. The report also presented all of the activities associated with each crop, the database cluster that would represent those activities for each crop, specific transfer coefficient values that should be used for each cluster, and a statistical examination of the structure of transfer coefficient data for each cluster. The final ARTF TC database contains information from 47 re-entry exposure studies, with a full set of body part transfer coefficients for over 750 worker-days. These data are organized into 28 transfer coefficient clusters that have a similar magnitude and distribution of exposure potential among body parts. The database also contains over 3000 C/A/H/F combinations assigned to those clusters to be represented by cluster-specific transfer coefficient data during risk assessments. In addition, the database identifies approximately 1400 C/A/H/F combinations that involve little or no contact between re-entry workers and foliage and for which the use of a transfer coefficient is not appropriate. Since all ARTF-conducted studies and some purchased studies used both inner and outer passive dosimetry (i.e., long underwear and outer clothing were analyzed following postapplication exposure), the TC database also contains a wealth of information on clothing penetration during re-entry activities (Baugher, 2006). A summary of the rationale for 28 clusters, and the descriptions of those clusters, follows.

53.7.2  Nine Special Clusters For various reasons, several crop situations are appropriately placed into their own unique cluster(s). Activities involving contact with turf are one such example since transfer coefficients obtained from transferable turf residues (TTRs) are not applicable to other crops for which dislodgeable foliar residues (DFRs) are an appropriate measure of the transferability of residues. In addition, golf course transfer coefficients were based on TTRs from undisturbed turf areas since some activities (i.e., mowing and hand watering) can affect TTRs on the course, so these transfer coefficients are not applicable to other turf situations. Based on these considerations, and a difference in the magnitude and distribution of transfer coefficients among body areas, the ARTF database now includes two turf clusters: DH: Sod: mechanical harvesting, scouting, transplanting, and hand weeding DM: Golf courses: maintenance activities. Greenhouse and nursery activities generally involved predominantly hand exposure and seem distinct from typical outdoor crop situations. Based on differences in total transfer coefficient values, three greenhouse-related ­ clusters were proposed:

Hayes’ Handbook of Pesticide Toxicology

GHf: Greenhouse floriculture hand harvesting: all flowers and methods GHv: Greenhouse vegetables: hand harvesting and similar contact activities GN: Greenhouse and nursery crops (others): all activities AND transplanting all crops. Irrigating field crops complicates the measurement of DFRs since the activity itself can impact the DFR measurement. Therefore, ARTF conducted a special study involving hand carrying and setup of irrigation lines, solid pipe sections with sprinkler heads that attach together in the field along rows to be irrigated. This study involved special DFR sampling since each DFR sample included leaf discs from a moist area, from which the pipes were taken, and a dry area, to which the pipes were carried. Since the potato plants were up to 3 feet tall, this study represents a worst case for irrigation. Most other irrigation methods (flood, ditch, overhead sprinkler, etc.) involve much less contact between worker and foliage. Therefore, ARTF proposed that the potato irrigation study represent irrigation in any crop where hand-line irrigation is commonly performed. All of the crops where growers or experts indicated irrigation occurs were designated as either involving or not involving hand-line irrigation. Those activities that do not involve hand-line irrigation were assigned to the No Transfer Coefficient cluster. One cluster is proposed for hand-line irrigation: I: Irrigation, any crop where hand-line irrigation is possible. Harvesting cotton is another special situation since the plant is desiccated, making collection of DFRs impossible. One study purchased by ARTF involved measuring dislodgeable residues on cotton bolls (DBRs) as well as dermal exposure following four activities associated with mechanical harvesting of cotton. Transfer coefficients for this crop would not be applicable to other crops where DFRs are more appropriate. Based on differences in magnitude and distribution of transfer coefficient measurements, three clusters for cotton harvesting were proposed: CHp: Cotton, mechanical harvesting: picker operator and raker (based on DBRs) CHm: Cotton, mechanical harvesting: module builder operator (based on DBRs) CHt: Cotton, mechanical harvesting: tramper (based on DBRs).

53.7.3  Eleven Field Crop Clusters The variety of crop profiles and re-entry activities within the field crop agronomic group is extensive. Examination of measured TC values indicated that leaf type appears to be an important determinant of transfer potential. Within each leaf type, crop height and/or the nature of the activity itself appears to determine the magnitude and distribution

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of exposure. Therefore, the ARTF database includes the following field crop clusters: Five smooth-leaf field crop clusters:

with differences in leaf surface characteristics. Therefore, the final cluster proposal for orchard crops is activity based and includes the following four clusters:

Sx: Smooth-leaf field crops: intense contact activities Ssr: Smooth-leaf field crops: scouting in row conditions Sss: Smooth-leaf field crops: scouting in solid stand conditions SH: Smooth-leaf field crops: hand harvesting and tying SW: Smooth-leaf field crops: hand weeding, thinning, and similar contact activities.

OH: Orchard crops: hand harvesting and similar contact activities OHn: Orchard crops: mechanically harvesting nuts OP: Orchard crops: hand pruning, scouting, and similar contact activities OW: Orchard crops: hand weeding and similar contact activities.

Three hairy-leaf field crop clusters: HH: Hairy-leaf field crops: hand harvesting and similar contact activities HHt: Hairy-leaf tobacco: hand harvesting and canopy management HS: Hairy-leaf field crops: scouting and similar contact activities. Three waxy-leaf field crop clusters: Wm: Waxy-leaf field crops, medium height: all activities, plus full foliage weeding WlH: Waxy-leaf field crops, low height: hand harvest and similar contact activities WlS: Waxy-leaf field crops, low height: scouting and similar contact activities.

53.7.4  Four Trellis Crop Clusters This agronomic group includes all the foliar contact activities identified by growers or experts that were performed on crops that are commonly grown on trellises. These crops (primarily grapes and caneberries) involve only smoothleaf crops, and none of the re-entry activities (except transplanting) are performed when the crop is of a low height. Therefore, clustering based on leaf type or crop height is not appropriate and the clustering is completely activity based. The following clusters reflect differences in transfer coefficient magnitude and distribution among body areas: Tx: Trellis crops: intense contact activities THg: Trellis crops: hand harvesting grapes and similar contact activities THb: Trellis crops: hand harvesting caneberries and similar contact activities TP: Trellis crops: hand pruning, scouting, and similar contact activities.

53.7.5  Four Orchard Crop Clusters This agronomic group involves activities only to high crops (except transplanting). Although there are transfer coefficient and DFR data for smoother (non-citrus) and waxier leaves (citrus), differences in the transfer coefficients measured for harvesting are not consistently associated

53.7.6  Transfer Coefficient Statistical Model With completion of the full database, ARTF has examined whether the exposure and DFR data supporting the transfer coefficients within each cluster are consistent with the original assumption of a proportional relationship between exposure and DFR. The analysis demonstrated that, for most clusters, the available data are consistent with the expected relationship between exposure and DFR. The simple proportional relationship of transfer coefficient  exposure/DFR is therefore a reasonable approximation. There are some cases in which the relationship cannot be discerned because exposure values were generated over a limited range of DFR levels. Although these clusters do not contain adequate data to support or counter the basic assumption of proportionality, the transfer coefficients nevertheless provide a reasonable measure of the ratio of exposure to DFR in these situations and may be applied to other situations under the assumption of proportionality that is shown will hold true elsewhere. In general, however, these cases are restricted to lower-exposure clusters that generally do not drive conclusions of product risk assessments. Therefore, the user of the ARTF Transfer Coefficient database can extract the appropriate transfer coefficient for the crop cluster of interest and, along with the appropriate DFR value, estimate the re-entry exposures for different work task activities. The DFR may be obtained either from a chemical-specific DFR study or by utilizing the EPA default assumption that the initial DFR is 20% of the application rate with a dissipation of 10% per day (U.S. EPA, 2000).

Conclusion The development of PHED in North America and UKPOEM and the German model within the European Union has provided powerful predictive tools for estimating worker exposure for specific pesticide use scenarios. Detailed knowledge of the workings of the database and models, appropriate selection of pesticide use information, and familiarity with the features and limitations

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of pesticide worker exposure studies are critical to their effective use. The unified North American and EU transition to AHED will provide a next-generation generic exposure database tool. The development of a worker reentry exposure database provides an important predictive tool for selected worker activity/crop combinations and, thus, fills an important gap for exposure assessors. Use of these widely available tools will continue to provide guidance to regulatory agencies and the agrochemical industry in regulation, product development, and product stewardship.

Acknowledgments The authors would like to dedicate this chapter to the memory of Joop van Hemmen, an inspiration to all of us and, more importantly, a dear friend.

References Agriculture and Agro-Industry including Fisheries (AIR). (1996). The Development, Maintenance, and Dissemination of a European Predictive Operator Exposure Model (EUROPOEM) Database. A EUROPOEM Database and Harmonised Model for Prediction of Operator Exposure to Plant Protection Products. Draft Final Report, December, Publication No. AIR3 CT93-1370, concerted action under the AIR specific programme of the Community’s Third Framework Programme for Research and Technological Development and managed by DGVI.FH3. Baugher, D.G. (2006). Penetration of Clothing by Dislodgeable Foliar Residues of Pesticides During Agricultural Occupational Reentry. ARTF, LLC. 22 February 2006. EPA MRID 46789302. Bruce, E.D. and Korpalski, S.J. (2008). Development of the ARTF Transfer Coefficient Database. Submission to the FIFRA Scientific Advisory Panel. OPP Regulatory Public Docket EPA-HQ-OPP-2008-0673. Curry, P. B., Iyengar, S., Maloney, P. A., and Maroni, M. (1995). “Methods of Pesticide Exposure Assessment,” Plenum, New York. Day, E. W. (1991). “Historical Development and Status of the Pesticide Users Exposure Database (A historical summary dated July 19, 1991),” Dow-Elanco, Indianapolis. Driver, J. H. and Whitmyre, G. K. (1997). Pesticide regulation and human health: The role of risk assessment. In “Fundamentals of Risk Analysis and Risk Management” (V. Molak ed.), pp. 143–162. CRC Lewis, New York. Ecobichon, D. J., ed. (1999). “Occupational Hazards of Pesticide Exposure: Sampling, Monitoring, Measuring,” Taylor & Francis, Philadelphia. Federal Biological Research Centre (FBRC). (1993). Guidelines for the Examination of Plant Protection Products in the Authorization Procedure. Part I, 3rd ed. Labeling of Plant Protection Products—Health Protection. Instructions for the Protection of Operators and Other Persons in the Directions for Use.” Braunschweig Federal Biological Research Centre for Agriculture and Forestry, Federal Republic of Germany. FIFRA Scientific Advisory Panel, (2007). Review of Worker Exposure Assessment Methods, SAP Minutes No. 2007–03, January 9–12, 2007. Franklin, C. A. and Worgan, J. P. (eds) (2005). “Occupational and Residential Exposure Assessments for Pesticides,” John Wiley & Sons, Ltd, West Sussex.

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Fong, H. R., and Krieger, R. I. (1988). Estimation of Exposure of Persons in California to Pesticide Products that Contain Dinocap (Karathane) and Estimation of Effectiveness of Exposure Reduction Measures. Publication No. HS-1469, Department of Pesticide Regulation, Worker Health and Safety Branch, Sacramento. Hackathorn, D. R. and Eberhart, D. C. (1985). Data-Base Proposal for use in predicting mixer-loader-applicator exposure. In R. C., Honeycutt, G. Zweig, and N. N. Ragsdale, (eds.), Dermal Exposure Related to Pesticide Use. ACS Symposium Series, Vol. 273. Am. Chem. Soc, Washington, DC, pp. 341–355. Hamey, P. Y. (1992). “Predictive Operator Exposure Model (POEM): A User’s Guide,”. MAFF Pesticides Safety Division. Honeycutt, R. C. (1985). Field worker exposure: the usefulness of estimates based on generic data. In R. C., Honeycutt, G. Zweig, and N. Ragsdale, (eds.), Dermal Exposure Related to Pesticide Use. ACS Symposium Series, Vol. 273. Am. Chem. Soc, Washington, DC, pp. 369–375. Honeycutt, R. C. and Day, E. W. Jr. (eds) (2001). “Worker Exposure to Agrochemicals: Methods for Monitoring and Assessment,” Lewis Publishers, Boca Raton. Honeycutt, R. C., Zweig, G., and Ragsdale, N. N. (eds.), (1985). “Dermal Exposure Related to Pesticide Use.” ACS Symposium Series, Vol. 273. Am. Chem. Soc, Washington, DC. Krieger, R., Blewett, C., Edmiston, S., Fong, H., Gibbons, D., Meinders, D., O’Connell, L., Ross, J., Schneider, F., Spencer, J., and Thongsinthusak, T. (1990). Gauging pesticide exposure of handlers (mixer/loaders/applicators) and harvesters in California agriculture. Med. Lav. 81(6), 474–479. Lunchick, C, Hamey, P., and Iyengar, S., (1994). The use of the North American (PHED) and United Kingdom (POEM) worker exposure models in pesticide registration. In “Proceedings of the Brighton Crop Protection Conference: Pests and Diseases–1994.” Vol. 3. British Crop Protection Council, Farnham, Surrey, UK. Lundehn, J. R., Westphal, D., Kieczka, H., Krebs, B., Locher-Bolz, S., Maasfeld, W., and Pick, E. D., (1992). “Uniform Principles for Safeguarding the Health of Applicators of Plant Protection Products: Uniform Principles for Operator Protection.” Braunschweig Federal Biological Research Centre for Agriculture and Forestry, Federal Republic of Germany. Maddy, K. T., Edmiston, S., and Richmond, D. (1990). Illnesses, injuries, and deaths from pesticide exposures in California, 1949–1988. Rev. Environ. Contam. Toxicol. 114, 57–123. Martin, A. D. (1986). “Estimation of Exposure and Absorption of Pesticides by Spray Operators.” Scientific Subcommittee on Pesticides and British Agro-chemical Association Joint Medical Panel Paper Number PS 4221/SC 8001. Martin, A. D., (1990). A predictive model for the assessment of dermal exposure to pesticides. In R. C., Scott, R. Guy, and J. Hadgraft (eds.), “Proceedings of a Workshop Entitled ‘Prediction of Percutaneous Penetration.” IBC Technical Services, London. Mehler, L., Thongsintusak, T., and Haskell, D., (1991). “Estimation of Exposure of Persons in California to Pesticide Products That Contain Benomyl.” Publication No. HS-1557, Department of Pesticide Regulation, Worker Health and Safety Branch, Sacramento. National Agricultural Chemicals Association (NACA). (1987). Jan. 2 letter from Dr. John McCarthy, NACA Director of Scientific Affairs to the NACA Exposure Assessment Subcommittee, regarding “Submission of data to the generic mixer/loader/applicator exposure database.” Nigg, H. N., Stamper, J. H., and Queen, R. M. (1984). The development and use of a universal model to predict tree crop harvester pesticide exposure. Am. Ind. Hyg. Assoc. J. 45, 182–186.

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Nutley, B. R. and Cocker, J. (1993). Biological monitoring of workers occupationally exposed to organophosphorous pesticides. Pestic. Sci. 38, 315–322. Pest Management Regulatory Agency (PMRA). (2002). “Canadian PHED Tables Final Version,” Operator Exposure Assessment Section, PMRA, Ottawa, ON. Plimmer, J. (ed.), (1982). “Pesticide Residues and Exposure.” ACS Symposium Series. Vol. 182. Am. Chem. Soc, Washington, DC. Rech, C., Bissell, S., and del Valle, M. (1988). “Potential Dermal and Respiratory Exposure to Abamectin during Greenhouse Applications.” Publication No. HS-1491, Department of Pesticide Regulation, Worker Health and Safety Branch, Sacramento. Reinert, J. C., and Severn, D. J., (1985). Dermal exposure to pesticides. The Environmental Protection Agency’s viewpoint. In R. C., Honeycutt, G., Zweig, and N. N. Ragsdale, (eds.), “Dermal Exposure Related to Pesticide Use.” ACS Symposium Series, Vol. 273. Am. Chem. Soc, Washington, DC, pp. 357–368. Saleh, M. A., Blancato, J. N., and Nauman, C. H. (eds.), (1994). “Biomarkers of Human Exposure to Pesticides.” ACS Symposium Series, Vol. 542. Am. Chem. Soc, Washington, DC. Thompson, R. (1998). “Agricultural Worker Crop Contact from Reentry Activities Performed in the USA and Canada: Grower Results. Doane Marketing Research, Inc. (USEPA MRID 44802601).” Thongsinthusak, T., Blewett, T. C., Ross, J., and Krieger, R. I. (1993). “Estimation of Exposure of Persons in California to Pesticide Products That Contain Chlorothalonil.” Publication No. HS-1475, California Department of Pesticide Regulation, Worker Health and Safety Branch (WHSB), Sacramento. U.S. Environmental Protection Agency (U.S. EPA) (1984). “Pesticide Assessment Guidelines. Subdivision K. Exposure: Reentry Protection,” Office of Pesticide Programs, Washington, DC.

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U.S. Environmental Protection Agency (U.S. EPA). (1995a). “Pesticide Handlers Exposure Database.” Occupational and Residential Exposure Branch, Office of Pesticide Programs, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA). (1995b). “Pesticide Handlers Exposure Database (PHED) Evaluation Guidance, PHED version 1.1.” Occupational and Residential Exposure Branch, Office of Pesticide Programs, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA). (1996). “Series 875—Occupational and Residential Exposure Test Guidelines, Group B—Post-Application Exposure Monitoring Test Guidelines.” Office of Pesticide Programs, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA). (1998). “PHED Surrogate Exposure Guide. Estimates of Worker Exposure from the Pesticide Handlers Exposure Database, version 1.1.” Office of Pesticide Programs, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA). (2000). “Policy Number: 003.1 Agricultural Transfer Coefficients.” Office of Pesticide Programs, Science Advisory Council for Exposure. 7 August 2000. van Hemmen, J. J. (1992). Agricultural pesticide exposure databases for risk assessment. Rev. Environ. Contam. Toxicol. 126, 1–85. Versar. 1991. Pesticide Handlers Exposure Database (PHED). “Data Entry Diskette User’s Guide, version 1.0.” Report prepared by Versar Inc., Springfield, VA for the U.S. EPA, Office of Pesticide Programs, Health Effects Division, Occupational and Residential Exposure Branch. Wang, R. G. M., Franklin, C. A., Honeycutt, R. C., and Reinert, J. (Eds.), (1989). “Biological Monitoring for Pesticide Exposures.” ACS Symposium Series, Vol. 382. Am. Chem. Soc, Washington, DC. Zweig, G., Leffingwell, J. T., and Popendorf, W. J. (1985). The relationship between dermal pesticide exposure by fruit harvesters and dislodgeable foliar residues. J. Environ. Sci. Health B 20(l), 27–59.

Chapter 54

Mitigation Measures for Exposure to Pesticides Thomas Thongsinthusak and Michael H. Dong Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, CA, USA

54.1  Introduction The most widely accepted paradigm for health or ecologic­al risk assessment is perhaps that presented by the U.S. National Research Council (NRC, 1983), which does not specifically include exposure mitigation as a key component. Yet even though the primary endpoint of such a framework is to develop an appreciation for potential risk, its ultimate goal is to alert and assist risk management in the mitigation of identified risk. The necessary inter­ action between risk assessment and risk management was not as apparent when the NRC paradigm was constructed (Goldstein, 2003), but became much more so upon further elaboration by the U.S. Environmental Protection Agency (EPA) (1995). Effective measures for exposure mitigation have been developed and/or advocated by many scientists involved in chemical exposure assessment, inasmuch as these individuals are typically the ones using these same measures as exposure factors and assumptions in their exposure evaluation. It is critical, therefore, that exposure assessors understand that while their primary responsibil­ ity is to perform exposure assessment, where warranted they should be ready to assist their risk management in the exposure mitigation by providing and advocating effective, practical mitigation measures. The NRC paradigm is actually composed of four key elements: hazard identification, dose–response assess­ ment, exposure assessment, and risk characterization (e.g., Faustman and Omenn, 2001). In essence, hazard identifi­ cation is typically the first step in health risk assessment undertaken by regulatory toxicologists to identify the adverse effects that can be induced by the toxic agent in question. Dose–response assessment is one further step undertaken to ascertain the identified cause and effect by analyzing the quantitative changes in effect caused by vari­ ous levels of exposure. As all toxicologists should agree, it is the dose that makes the poison. This dose–response Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

notion advocated by Paracelsus (1493–1541), who is often regarded as the father of modern toxicology, simply assures that no matter how toxic an agent is, the health risk in ques­ tion is significant where only the exposure is sufficient. It is based on this notion that exposure assessment plays an important role in health risk assessment. This third com­ ponent involves the systematic determination as well as estimation of the type, magnitude, duration, and frequency of exposure experienced or anticipated to occur. Normally, risk characterization is the final phase undertaken by regu­ latory toxicologists to integrate and characterize the analy­ ses performed in the other three components. It is expected that, soon after a health risk is charac­ terized under the NRC paradigm or a similar framework, health risk management would proceed with the necessary management or regulatory actions if the identified risk is determined to be significant. Although not much of this risk management function is described in the so-called red book that details the NRC paradigm in great length, such a mitigation process is frequently a regulatory or manage­ ment routine. In practice, regulatory scientists, especially those work­ ing at the exposure assessment end, are the ones that typi­ cally assist the risk management team in using exposure mitigation measures (EMMs) to reduce the exposure of concern. These scientists are the ones in the best position to judge scientifically whether or not a certain mitigation option is quantitatively as well as qualitatively sound. After all, these individuals are the ones responsible for characteriz­ ing the exposure scenarios for which the acute, intermediate, and long-term effects are assessed and, as mentioned earlier, are also the ones relying on the same measures as exposure factors and assumptions in their exposure assessment. EMMs are simply controls, methods, tools, restric­ tions, or otherwise actions taken by responsible author­ ities to (help) reduce the exposure(s) of concern. Over the years, researchers and regulators have proposed numerous 1157

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measures for the effective use of EMMs for children (Liebman et al., 2007) and for pesticide handlers/users (Fenner-Crisp, 2001; Grieshop, 1988; Krieger et al., 1992; Palis et al., 2006; Thongsinthusak and Frank, 2007; U.S. EPA, 1999a,b). Mitigation options available for regulatory use today generally include, but are not limited to, personal protective equipment (PPE), administrative control, engin­ eering control, application rate or usage, timing of appli­ cation, and cancellation of use. Some of these options can be employed as replacements of certain measures specified on the initial product labels. For example, according to the Worker Protection Standard (U.S. EPA, 2005a), pesticide handlers need not use all the PPE specified on a product label when certain engineering controls (e.g., an enclosed cab) are used. This federal standard thus provides certain flexibility for the application of effective EMMs. This chapter is an attempt to describe how EMMs can be developed and advocated more effectively by expos­ure assessors to assist their risk management in reducing the exposure(s) of concern. Due to the extensive use of pesti­ cides by the agricultural community, and the limited space available, this chapter focuses primarily on those EMMs used for the protection of agricultural workers (e.g., han­ dlers, users, fieldworkers) and of individuals who may come in contact with pesticide residues inadvertently (e.g., residents and bystanders). The discussion is further limited to exposures via the dermal and inhalation routes, given that dietary exposure is a topic unto itself. Another subtle reason is that dermal exposure typically accounts for over 90% of the combined daily exposure to most any pesticide used (Durham and Wolfe, 1962; Wolfe, 1976). One note­ worthy exception, however, is that for highly volatile pesti­ cides (such as methyl bromide and methyl isothiocyanate, which are commonly applied as soil fumigants) inhalation becomes the major route of exposure. This chapter begins by reviewing the basics of a pesticide EMM process, then considers the major technical issues inherent in the develop­ ment of the more feasible, more specific EMMs. Finally, for reference as well as illustration purposes, it concludes with several actual cases of pesticide EMMs used in the United States, including those used specifically in the state of California.

54.2  Basics of a pesticide EMM process Basic prerequisites for the development of pesticide EMMs, especially for those in California, are as follows: (1) The pesticide of concern typically should be registered and evaluated for potential adverse health effects; (2) the assessment process is supported or mandated by regula­ tions; (3) equipment or tools used for exposure mitigation are available at a reasonable cost; (4) handlers/users under­ stand that the pesticide can cause adverse health effects;

Hayes’ Handbook of Pesticide Toxicology

and (5) manpower is available to perform the mitigation tasks. For example, pesticides that are sold and used in the United States must first be registered with, and eventually be evaluated by, the U.S. EPA, as mandated by the Federal Insecticide, Fungicide, and Rodenticide Act of 1972. Moreover, the same pesticides must be additionally regis­ tered with, and evaluated by, the California Department of Pesticide Regulation (CDPR) before these products can be sold or used in the state of California. For the case in California, there are several additional or unique legislative mandates concerning the protection of human and environmental health, including notably the Birth Defect Prevention Act (BDPA) of 1984, which is explicitly specified in the California Food and Agricultural Code (CFAC-BDPA, 1984). The purpose of this important legislation is to prevent pesticide-induced abortion, birth defects, and infertility. Also included in the CFAC, Sections 12980–12988 (Article 10.5), is another equally important legislation entitled Pesticides and Worker Safety, which specifically stipulates that “It is necessary and desirable to provide for the safe use of pesticides and for safe working conditions for farmworkers, pest control applicators, and other persons handling, storing, or applying pesti­cides, or working in and about pesticide-treated areas.” This stipula­ tion assures CDPR a key role in reducing worker and resi­ dential exposures to an acceptable regulatory target level (benchmark). It is also based on this stipulation that CDPR is required to perform health risk assessments for priori­ tized pesticides that are sold and used in California. It is noteworthy here that across countries at the global level, pesticide handlers/users may have different perspec­ tives on health effects caused by pesticides. For instance, some pesticide handlers/users in the Philippines may think that pesticide exposure is only via inhalation and dietary and that they are immune to much of the toxic effects caused by pesticides (Palis et al., 2006). On the other hand, some pesticide handlers/users in Ecuador may believe that PPE (e.g., respirator, gloves) is uncomfortable to wear as well as too costly to maintain (Grieshop, 1988). These perspectives are different from those in the United States. Therefore, EMMs used for mitigation of pesticide expo­ sure are expected to vary considerably across countries. Further discussion on the relevant practicalities or imprac­ ticalities, particularly at the local level, is given in Section 54.2.2 (especially in 54.2.2.3). It is also interesting to note that the risk assessment process undertaken at CDPR is somewhat diff­erent from what has been advocated in the NRC paradigm. At CDPR, when results of a pesticide risk assessment perf­ormed under BDPA or any other regulatory mandate indicate an unacceptable risk, the relevant exposure(s) must be reduced before the pesticide under evaluation is allowed for con­ tinued registration. In California, EMMs are typically required when the ratio of the selected no observed effect level (NOEL) to the estimated exposure level, otherwise

Chapter | 54  Mitigation Measures for Exposure to Pesticides

known as margin of exposure (MOE), is determined to be substantially less than the benchmark of 100 for non­ cancer risk where the toxic endpoint of concern is derived from animal studies, or less than 10 if the toxic endpoint is observed on human subjects. For cancer risk involv­ ing agricultural workers, the general benchmark is set at 1  105 excess risk. Due partly to the inconsistency with the NRC scheme, but mostly in an effort to advocate a more effec­ tive developmental process for pesticide EMMs, CDPR (Gosselin, 2001) eventually issued an executive policy for guiding both the risk management and the risk miti­ gation processes. Under this policy, both the Pesticide Enforcement Branch at CDPR and the County Agricultural Commissioner Offices at the local level, and not the pes­ ticide risk or exposure assessors, have responsibility for enforcing the pesticide EMMs that are in effect.

54.2.1  Initiation of EMMs The time to initiate the EMM process for agricultural workers, where warranted, may be different from agency to agency or from country to country. For example, when the U.S. EPA issues its Reregistration Eligibility Decision (RED), RED Fact Sheet, or RED Facts for pesticides, such as for s-ethyl dipropylthiocarbamate (EPTC; U.S. EPA, 1999a), chlorothalonil (U.S. EPA, 1999b), and methyl bromide (U.S. EPA, 2008a), the document typically has already contained one or more sets of proposed EMMs. The general public and pesticide registrants can then pro­ vide comments on the RED or a similar document during the comment period. On the contrary, the EMMs proposed at CDPR are not typically included in its risk character­ ization or the accompanying exposure assessment docu­ ment even when the pesticide risks are determined to be un­acceptable. Only when these two documents are final­ ized and signed off would the management at CDPR issue a directive to alert its exposure mitigation unit (and other related units) that certain EMMs may be needed to reduce the pesticide exposures at issue.

54.2.2  Important Factors and Major Concerns One key concern for developing a feasible, specific pesti­ cide EMM is that the same PPE and engineering controls that are suitable for use in one region may not be appro­ priate or applicable in another, a point touched on earlier. For example, a chemical- or water-resistant coverall (e.g., rainsuit) provides an excellent barrier to chemicals from contacting the skin. However, the PPE can retain mois­ ture and trap heat from the worker’s body. When used under hot and humid conditions, this PPE thus can cause heatstroke or other undesired health effects. For this

1159

reason, the California Code of Regulations (CCR, 2008a) specifies that If the ambient temperature exceeds 80°F during daylight hours or 85°F during nighttime hours (sunset to sunrise), pesticides requiring a chemical-resistant suit are not handled by employees unless (either) they are handled pursuant to exceptions and substitutions permitted in (i) or employees use cooled chemical-resistant suits or other control methods to maintain an effective working environment at or below 80°F during daylight hours or 85°F during nighttime hours (sunset to sunrise).

In short, it is crucial for regulators to consider not only the mitigation benefits of using a particular PPE item, but also the health effects that it may cause. Some PPE items can be very costly to purchase and/or to maintain (including the replacement parts for decontamination). Furthermore, certain disposable PPE may be considered suitable in some regions or countries, but not necessarily a good option in some others due to different local environ­ mental laws or regulations governing waste disposal. Engineering controls used as pesticide EMMs typically include the use of an enclosed cab with or without positive pressure and a closed mixing/loading system. These and some other engineering controls are excellent measures for use to reduce dermal and inhalation exposures. However, in addition to their high cost, proper training is required for workers to use these controls effectively, since they are available in numerous variations and configurations. Subchapter 3 of CCR (2008b) contains extensive infor­ mation on pesticide worker safety in California. It prescribes proper training for handlers, medical supervision, change areas, handler decontamination facilities, PPE, respiratory protection, equipment maintenance, closed systems, proper training for fieldworkers, fieldworker decontamination facilities, and restricted entry intervals (REIs). Under these state regulations, CDPR is required to ensure that agricul­ tural workers use pesticides in pest control operations in a proper manner, that these workers are adequately protected, and that their illnesses/injuries are minimized from using pesticides or from working in crops sprayed with pesti­ cides. Naturally, different countries would have different regulations governing their own pesticide use for pest con­ trol. Accordingly, the methods and programs used in health protection for workers in other places are not expected or needed to be the same as those used in California, so long as the workers are adequately protected.

54.2.2.1  Protection Factors Used in California Unless chemical-specific protection factors are available, default protection factors (DPFs) are usually employed in developing EMMs for most pesticides with a bias erring on the side of worker and public health protection. DPFs are those factors determined mostly from field studies covering collectively a few to several pesticides. The DPFs used at

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CDPR were obtained from the open literature, field stud­ ies submitted by pesticide registrants, and field studies conducted by CDPR. These defaults have been published in several places in the literature (e.g., Aprea et al., 1994; Gerritsen-Ebben et al., 2007; Thongsinthusak and Ross, 1994; Thongsinthusak et al., 1991, 1993; U.S. EPA, 1998a). Table 54.1 summarizes the interim DPFs adopted by CDPR for various EMMs used in reducing pesticide exposure. Note that the EMM (developmental) process is intended to reduce pesticide exposures of individuals down to or below an acceptable regulatory target level. Therefore, even when these DPFs are each affixed to a given measure, users of these defaults should have the appreciation that the fabrics and other materials used to construct the PPE and non-PPE attires may indeed differ considerably across countries. And so the qualities and hence the effective­ ness of the various respirators and engineering controls are available across countries.

54.2.2.2  Protection Factors Used in Other Regions More recently, Gerritsen-Ebben et al. (2007) in the Netherlands published a report summarizing the various DPFs adopted across regions. The report was a result of literature reviews of the inherent issues considered by vari­ ous regulatory authorities in North America, Europe, and Australia, as well as by several industry organizations and academic groups working on the issues in these regions. The Netherlands investigators noticed some variance in approaches between the North American and European authorities (see Table 54.2) and hence proposed a number of DPFs for harmonized use. Note that although a uniform application of these harmonized DPFs may be feasible in the future, it may still take some time before those regula­ tory authorities can achieve consensus.

54.2.2.3  Practicalities and Limitations at the Local Level Over the years, numerous types of EMMs in various forms, including product cancellation, have been considered and implemented quite successfully. Yet not all EMMs in use today are applicable to all people working or living in the same general area, much less across regions or countries, as pointed out in earlier sections. Some mitigation options tend to be applicable to handlers or users alone, whereas others are unique to fieldworkers or to residents/bystanders. It is also important to realize that not all applicable or unique measures are equally feasible or effective in all settings, even when they are applied to the same group of people. The practicalities of these mitigation measures at the local level, much like at the global level, are each like­ wise bound by many factors including those that are socio­ economic, health-based, and/or political in nature.

Hayes’ Handbook of Pesticide Toxicology

For example, as further elaborated on in Section 54.3.2, the single major EMM for fieldworkers is (an extension of) REI or pre-harvest interval (PHI), either of which is considered to be a highly practical administrative control. At times, an altered application method or reduced prod­ uct usage may be employed as an EMM, though often not regarded as equally effective or practical. Rarely is PPE or additional clothing considered to be appropriate for fieldworkers, since many of their reentry activities (e.g., harvesting and canopy management) are extremely labor intensive and require a high degree of mobility and dex­ terity. As alluded to in Section 54.2.2, the health effects of heat stress from putting on additional clothing or PPE cannot be overlooked. In addition, many fieldworkers are paid by the amount or the volume of work performed. For this reason, they tend to put on as few pieces of clothing as practical when working in a field. Requiring PPE or reducing product usage could com­ promise agricultural productivity while at the same time the mitigation in exposure might not be sufficient. This leads to the general consensus that the use of an extended REI would be the most practical, if not the most effective, mitigation option for fieldworkers. It is also important to note that even though PPE and additional clothing may be considered more feasible EMMs for handlers/users, the proper maintenance and supply of these gear are always a practical issue that may be overlooked by those advocating such use. Many EMMs are known to be unenforceable when used on residential property. Additionally, the pesticide exposures of sensitive populations, for example, children and the elderly, may require larger MOEs. One regulatory option commonly used for residential exposure is product cancellation. Short of this, one relatively workable miti­ gation measure is through use of more comprehensible product labeling or label specifications, along with safer packaging, safer application apparatus, or improved for­ mulation technologies (e.g., microencapsulation). Another way to help mitigate the health concern for residents is through health education and/or health risk communica­ tion, which may be political, socioeconomic, or public health-based actions. For bystanders and nonuser residents, a posting require­ ment can be one of the most practical administrative con­ trols. These individuals need to be informed about the location and timing of the pesticide application to ensure that they do not enter areas treated recently. This require­ ment is especially important for use of buffer zones as an EMM. Bystanders including those working in adjacent fields, not just nonuser residents living nearby, need to be given such information so that they will not enter areas designated as a buffer zone. A buffer zone provides dis­ tance between the application site (e.g., field edge) and bystanders, tending to allow airborne residues to disperse before reaching the bystanders. This buffer will reduce the

Chapter | 54  Mitigation Measures for Exposure to Pesticides

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Table 54.1  Default Protection Factors for Various Mitigation Measures Applied to Handler Exposurea Protective item [body region protected]

Protection factor (% exposure)b

A. Personal protective equipment   Coveralls [all but head, hands, feet]

90

  Chemical-resistant full-body protective clothing [all but head, hands, feet]

95

  Chemical-resistant apron [chest/stomach, front half of thighs]

95

  Chemical-resistant gloves [hands]

90c

  Goggles, nonvented [½ of face, or ¼ of head]

95d

  Goggles, vented [½ of face, or ¼ of head]

75e

  Face shield [face]

75

  Chemical-resistant boots [feet]

90f

B. Nonpersonal protective attire  

Short pants and a short-sleeved shirt [chest/stomach, back, upper arms, thighs]

90g

 

Long pants and a long-sleeved shirt [all but head, neck, hands, feet]

90g

 

Shoes plus socks [feet]

90

C. Air-purifying respirators: particulate and gas/vapor filters   Dust/mist filtering respirator, approved by MSHAh/NIOSHi

90j

 

Half-face, approved by MSHA/NIOSH

90j

 

Full-face, approved by MSHA/NIOSH or a NIOSH-approved respirator with appropriate N, R, P, or HE prefilterk

99j

D. Powered-air purifying respirators (PAPRs): dust/mist/fume filters   Half-mask, approved by MSHA/NIOSH

98j

  Full-face or helmet/hood, approved by MSHA/NIOSH

99j

E. Supplied-air respirators (SARs) or airline respirators  

Full-face and operated in a pressure-demand or other positive-pressure mode, approved by MSHA/NIOSH

99.9j

F. Self-contained breathing apparatus (SCBA)  Full-face or other full sealing system and operated in a pressure-demand or other positive pressure mode, approved by MSHA/NIOSH

99.99j

G. Engineering controls   Closed mixing/loading system

95l

  Enclosed cab with positive pressure and charcoal air filtration unit meeting S525 standardm

98

  Enclosed cab

90

a Adapted primarily from Thongsinthusak and Frank (2007), courtesy of ACS (American Chemical Society Symposium Series 951), which was actually based on the work of Thongsinthusak et al. (1991, 1993); additional respiratory protection factors were from ANSI (2006); for chemical-resistant apron, goggles, shoes plus socks, chemical-resistant boots, and face shield, they were established by the Human Health Assessment Program at CDPR during their staff meetings. b Protection factors (PF, %) for personal protective equipment (PPE) and clothing are applied to dermal exposure of the specified body regions only; PF for respirators and enclosed cabs are applied to inhalation exposure; and PF for closed systems is applied to both dermal and inhalation exposure. c Aprea et al. (1994). d Based on the assumption that goggle material provides protection similar to that of chemical-resistant apron or suit. e Based on the assumption that aerosols and airborne residues can pass through openings. f Based on the assumption that chemical-resistant boots give the same protection as from use of chemical-resistant gloves. g Based on the assumption that the protection is similar to that from use of coveralls. h MSHA  Mine Safety and Health Administration. i NIOSH  National Institute for Occupational Safety and Health. j American National Standard for Respiratory Protection Z88.2 (ANSI, 2006). k U.S. EPA (1998a). The new designations for filters N (not resistant to oil), R (oil resistant), or P (oil proof) each have three levels of filter efficiency: 95, 99, or 99.97% (with the latter sometimes being loosely considered as 100%). HE (high-efficiency particulate aerosol) is a category of canister-type respirators with a particle filtering efficiency above 99.97%. HE respirators are rarely used and are more expensive than those in the 95%, 99%, and 99.97% particle removal efficiency categories. Filter efficiency typically employs the most penetrating aerosol size of 0.3 mm aerodynamic mass median diameter (CFR, 2008). l Thongsinthusak and Ross (1994). m ASAE (American Society of Agricultural Engineers) Standard S525 (1998a,b).

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Table 54.2  Default Protection Factors Used by Various Regulatory Authoritiesa,b Mitigation measurec

CDPR

U.S. EPA

PMRA

PSD

Overalla

  Clothingd

90

50

75

—

90

  Coated coveralls

95

—

90

  Gloves

90

A. PPE/non-PPE for handlers

90 90–99

90/95e

B. PPE for fieldworkersf   Clothingd

—

—

—

—

80

 Closed system for mixing/ loading

95–98

90–98

—

—

90/95g

 

90

—

—

—

90

C. Engineering controls

Enclosed cab

a

From Gerritsen-Ebben et al. (2007), as percent of dermal and/or inhalation exposure, where applicable. CDPR  California Department of Pesticide Regulation; U.S. EPA  U.S. Environmental Protection Agency; PMRA  Pest Management Regulatory Agency of Health Canada; PSD  Pesticide Safety Directorate of the United Kingdom (and the European Union). c PPE  personal protective equipment; those measures not listed here (e.g., chemical-resistant apron, chemical-resistant boots, goggles) were not considered in the report by Gerritsen-Ebben et al. (2007). d Including long pants and a long-sleeved shirt (cotton/cotton-polyester) and uncoated coveralls. e 90% and 95% have been proposed for gloves where liquids and solids are handled, respectively. f Only limited data were available for this group of workers. g 90% and 95% have been proposed for closed systems where liquids and solids are handled, respectively. b

chances that the air concentrations will cause acute adverse health effects in places where bystanders are located. Currently, the U.S. EPA is requiring fumigant users to establish a buffer zone around treated areas to help reduce risks from acute inhalation exposure of bystanders. Further elaboration on the EMMs for residents and bystanders is given in Section 54.3.3.

54.3  Criteria for more practical, more specific EMMs When public agencies are required to protect the health and safety of individuals, particularly the agricultural workers as mandated by environmental health laws and regulations, the regulators or their scientists should be able to guide the mitigation measures effectively for the pesti­ cide exposure(s) of concern. That is, these individuals should have a fair amount of knowledge about the applica­ tion method, formulation, crop/site under treatment, expos­ ure period, and route of exposure that alone or in some combination would yield the highest anticipated exposure level (i.e., the lowest calculated MOE). Otherwise, they would be unable to propose reasonable EMM(s) in such a way that the highest anticipated exposure level would be reduced to an acceptable target level. Other criteria considerations for the development of an effective pesticide EMM include, but certainly are not limited to, the subpopulation affected or to be affected by

the pesticide exposure at issue; the cost and availability of the mitigation option used; and the climatic, land, or other environmental conditions under which the EMM is imple­ mented. Despite the fact that many EMMs are available to reduce pesticide exposures considerably and effectively, not all workers or growers can comply with all these miti­ gation requirements, since high compliance cost and/or practical limitations may be involved. For example, chemicalresistant PPE (e.g., coveralls) can effectively reduce pesti­ cide exposure but may not be suitable for use in some areas where the temperature and humidity are unusually high. As discussed in Section 54.2.2.3, when certain pesti­cide EMMs are considered suitable for use in one local area, there is no assurance that the same are applicable in another local area, sometimes within the same general agricultural vicinity for different work activities.

54.3.1  EMMs for Pesticide Handlers/Users Where necessary and suitable, several mitigation options can be used in combination to reduce handler (herein includ­ ing user) exposures to an acceptable regulatory benchmark. As evident from the DPF data in Table 54.1, the options are PPE (e.g., clothing or respirators) and engin­eering controls (including for example use of an enclosed cab, a closed mixing/loading system, a piece of improved application equipment, or an improved formulation such as watersoluble bags). Note that PPE requires user compliance to be

Chapter | 54  Mitigation Measures for Exposure to Pesticides

effective although it may be less expensive, whereas engin­ eering controls are typically more expensive and passive. Again, the DPFs for these mitigation options are listed in Table 54.1 for use specifically in California, and in Table 54.2 for use in additional places. Depending on the pesticide used, other options may include limited work hours per workday or workweek, timing of application during the day or the season, designation of the active ingredient (AI) as a minimal exposure pesticide (California only) or as a restricted use pesticide, and reduction of application rate or pesticide usage. If the exposures to a pesticide AI cannot be reduced to the acceptable levels via these measures, cancel­ lation of use may become the final option. Any mitigation option proposed for the pesticide AI under evaluation will eventually need to be reviewed by interested parties, typically the environmental groups and the pesticide industry. Inputs from both the general pub­ lic and industry are considered to be crucial in that these EMMs must be practical to pesticide handlers while at the same time health protective to these workers and the general public. To this end, acceptable target levels of exposure should be established beforehand, so that these benchmarks can be used to gauge or steer the proper course for adopting the most effective EMMs. There are at least two distinct regulatory mechanisms through which mitigation options can be implemented, especially at the state or local level. These regulatory mechanisms were in fact used for EPTC and methyl bro­ mide handlers in California, which are further discussed in Sections 54.4.2.1 and 54.4.2.2 (respectively) for illus­ tration purposes. In essence, the EMMs used for EPTC in California have been attached to the federal product labels, specifying that the mitigation required is for California handlers only. And those used for methyl bromide handlers in California, are being implemented by means of state regulations.

54.3.2  EMMs for Fieldworkers As alluded to in the introduction, the effective use of pesticide EMMs, whether for fieldworkers or for other groups, should begin with an understanding of the way in which pesticide exposures are identified and assessed. Accordingly, a brief discussion of the basic assessment process is warranted here for field reentry exposure, including how REIs are determined. In pesticide exposure assessment, field reentry exposure is defined as a person’s exposure to a pesticide from enter­ ing, staying, or working in fields previously treated with that pesticide. The individuals considered are those work­ ing in the agricultural fields and, hence, are often referred to as fieldworkers. Because in all cases fieldworkers are not allowed to enter a treated field until sprays have dried and dust settled, the airborne residues are typically negli­ gible, and dermal contact should be much more significant

1163

than contact through inhalation. One general observation supporting this consideration is that most pesticides have a low vapor pressure. Pesticides of low vapor pressure tend to be more persistent and hence more available on foliage where pests harbor. The assessment of field reentry exposure to pesticides, or of reentry exposure for short, rests on two key compo­ nents: (1) the amount of pesticide available for contact and (2) the way in which the contact is made. Conceptual and technical issues associated with the two components are both complex and manifold, which are the focal points of discussion in this section. In practice, reentry exposure is estimated using the sim­ plified algorithm as follows:



[ Reentry dermal exposure]  [amount available for dermal contact ]  [dermal transfer rate ]

(1)

where the term dermal transfer rate (TR) represents the overall joint transferability effects of all the inherent fac­ tors and subfactors involved in assessing both the amount of residues available for contact and the way in which a contact is made.

54.3.2.1  Calculation Algorithm for Reentry Exposure In one form or another, Eq. (1) has been discussed exten­ sively in the literature (e.g., Brouwer et al., 2000; Krebs, et al., 2000; Popendorf, 1985; U.S. EPA, 2000; van Hemmen et al., 2001). By convention and more for simplicity, the third term TR is often defined quantitatively as the ratio of dermal residues in g/h of reentry task to the underlying dislodgeable foliar residues (DFRs) in g/cm2. Its value for a given reentry task is experimentally determined from some trial events. In calculating the TR value for cotton scouts, Dong (1990) provided an example in which dermal residues can be quantitatively related to DFR by means of a linear regression, thus not necessarily as a ratio. Note that in reality, the algorithm for approximation of reentry dermal exposure (RDE) is



 RDE trial   RDE current  DFR current    DFR trial 

(2)

where RDEtrial/DFRtrial is the ratio or mathematical (sta­ tistical) relation that sets the TR value (Dong, 2004). The TR value in Eq. (1) or (2) for a specific reentry task thus may be learned from one or more trials in which pesti­ cide residues on both the foliage and the worker’s body or clothing would be monitored concurrently to provide the ratio (or, rarely, the regression in some form). The term TR is most commonly found in the literature as transfer coeffi­ cient (TC). Where chemical-, crop-, and task-specific trial

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1164

data are unavailable, the TR value for a given reentry task is often approximated using those established for similar activities. Such a surrogate approach requires some scien­ tific judgment and necessarily a fair amount of conserva­ tism to err on the side of health protection. The term DFR is traditionally defined as the amount of pesticide residues that can be removed off both sides of treated foliage surfaces through use of some standardized aqueous surfactant and mechanical agitation in an ana­ lytical laboratory (Iwata et al., 1977). Guidance has been provided regarding the statistical (e.g., Andrews, 2000) and the laboratory (e.g., del Valle, 1999) analysis of DFR and the procedures for collection of foliar residue samples (e.g., Edmiston et al., 2002). While by definition the DFR values serve much like an environmental index, they are rather useful tools for esti­ mation of reentry exposure. When coupled (multiplied) with a TR, this index can be readily converted to hourly (and eventually daily) dermal exposure of fieldworkers reentering a treated field. Conceptually, DFR levels are used like the temperature of a thermometer. Where the DFR level is high, it suggests that the reentry exposure in question is likewise high or, in some instances, even unsafe. Yet the same DFR level may have different impli­ cations of reentry safety for different reentry activities and hence to different fieldworker groups.

54.3.2.2  Variability and Uncertainty The joint use of TR and DFR for estimation of reentry exposure is not without uncertainties. These uncertainties are inherent in the two algorithm components in Eq. (1) that are fundamental to all forms of human exposure assess­ ment, including handler exposure. These uncertainties are indeed real and manifold, and have strong effects on the REI. Some sources of uncertainty such as the influence of leaf surface features (waxy, smooth, hairy) on expos­ ure have been acknowledged, as detailed in Chapter 53 on exposure databases. The first component in Eq. (1) appears to be more intu­ itive in that if more is available for contact, then naturally there is more to be contacted. However, the rate of residue decay may not be described by a simple first-order decay process (Whitmyre et al., 2003), and the rate of decay for a particular crop and chemical may vary with season (Dong et al., 1992). The second component is much more com­ plex, even in concept. Many factors must be considered when assessing the way a contact is made. For example, it is very difficult to quantify either the intensity or the fre­ quency of contact, much less their joint effects, since here the contact is highly a human act. The preceding and other related issues collectively form the basis for the variability and the uncertainties that could have major effects on the REI used as an EMM for fieldworkers. Note that some of these same questions (and

hence the associated limitations and uncertainties) may be pertinent to handler exposure as ���������� well���.

54.3.2.3  Use of Extended REI as EMMs In spite of the limitations and uncertainties involved, Eq. (1) asserts that an extended REI still can be a practical, if not highly effective, EMM for fieldworkers. The algorithm further asserts that certain information on the DFR’s dis­ sipation behavior is central to the assessment of reentry exposure (and hence to its mitigation) insofar as the TR, the other component in the algorithm, is predetermined and hence is no longer a variable. Accordingly, foliar dissipa­ tion curves are often utilized to establish an REI for use to assure a DFR level within the safe region. It is important to note that even if the values of DFR were affordably measured every day within the exposure interval and thus available for a long enough time period, their dissipation behavior, such as half-life, should still be best expressed as (or simplified into) a mathematical equation. Otherwise, a fairly lengthy look-up chart might be needed in order to assign DFR values for REIs that are many days from application time. Also, logistically it is not feasible to monitor DFR every day in an exposure period, particularly when the foliar residues have a half-life longer than a few days. It is thus necessary to have an effective way to project the DFR level for any REI in question, inso­ far as the more practical measure for mitigation of reentry exposure is by imposing the shortest safe REI. As fur­ ther discussed in Section 54.4.2, foliar dissipation curves have been used as an actual exposure mitigation tool for establishment of the shortest REI that is still deemed to be sufficient to reduce DFR to a safe level. As with many other events in (pesticide) exposure assessment, the determination of foliar dissipation is both an art and a science. Yet, as referenced earlier, in its sim­ plest term the foliar dissipation of many pesticide residues can be described with fair accuracy by the first-order expo­ nential decay process, or by its equivalence, the log-linear regression, as follows (e.g., Andrews, 2000; Dong et al., 1992; Willis and McDowell, 1987):



DFR t  (DFR 0 )ekt [alternatively, log DFR(t ) (3)  log DFR(0)  (k )t ]

where DFR0  initial deposition, DFRt  DFR at time t (postapplication), and k  the first-order reaction constant. Whatever mathematical model is adopted, the highest DFRt level that still leads to an acceptable (dermal) re­entry exposure can be projected by selecting the shortest value for time t that still ensures safe reentry. This specific time t then becomes the shortest safe REI in question. The limita­ tions and uncertainties revolving around the use of REI for mitigation purposes are essentially the same as those dis­ cussed earlier regarding the use of Eq. (1) or (2).

Chapter | 54  Mitigation Measures for Exposure to Pesticides

54.3.2.4  Other Potential Mitigation Measures It is important to understand that the TR used for estimat­ ing reentry exposure should be based on dermal exposure monitored in trials where fieldworkers do not use any more PPE or clothing than required by the pesticide prod­ uct label. Otherwise, the reentry exposure to be calculated would be compromised (underestimated). As most regis­ tered pesticide product labels do not specify any PPE or specific clothing, the use of gloves, waterproof boots, or a long-sleeved shirt potentially can be an effective (though often much less practical) EMM for these workers. There are other considerations or issues of similar importance here. For example, although most hand har­ vesters prefer not to wear gloves while working in a field, many fruit tree pruners wear gloves to protect their hands from physical hazards. As shown in Table 54.1, gloves do have a very large dermal protection factor of 90% or higher. Yet, at the same time, one must be aware that there are situations in which the hand is not the predominant body region coming in contact with foliar residues. A case in point is for cotton scouts whose entire body sometimes must walk through rather tall and thick foliage to inspect treatment efficacy. Another example is when an irrigation worker may only need to walk over a low row crop to carry irrigation pipes around. In fact, for this type of irrigation work, waterproof boots may serve as an effective adjunct EMM to waterproof gloves. There are still other cases in which the use of TR should rely on the expectation that most, if not all, field­ workers would wear gloves in a field, regardless of label specification. For instance, it is more than common prac­ tice that harvesters cutting roses would wear gauntlet-type gloves to protect their hands and forearms, as roses are full of thorns. These workers typically put the cut roses on their forearms until the bundle that they are holding becomes large enough. If fieldworkers must make an early entry into a field treated with a fumigant or an otherwise highly volatile pesticide, then the use of a NIOSH-approved respirator may be the most viable EMM. As shown in Table 54.1, an acceptable respirator generally has an inhalation protection factor of 90% or higher.

54.3.3  EMMs for Residents and Bystanders Only very few EMMs are practical for residents and bystanders. This is expected given that most EMMs devel­ oped for workers are not enforceable when they are applied in a residential setting. People are not obligated, and many times are reluctant, to use certain PPE or clothing while they are on a residential property where a pesticide has been applied. Even homeowner users sometimes tend not to follow well-written instructions when applying pesti­ cides in or around their home.

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One effective EMM for residential exposure is to refor­ mulate the pesticide or redesign the product packaging so that the user or nonuser would receive less exposure when or after the product is applied on a residential property. For example, baits can be made available in a fully child-resistant or tamper-resistant container. Another health concern over residential exposure is that certain flea collars can be used continuously throughout the year. There does not seem to be any practical or enforceable measure available to miti­ gate such chronic exposure for children with treated pets. In many of these cases the only option may be to cancel the product’s registration, or alternatively to develop data dem­ onstrating negligible exposure. Another reality is that many herbicides are now allowed to be used on home lawns, whether by home­owners or commercial lawn care specialists. The most common label restriction is that residents are instructed not to be present on the treated lawn until sprays have dried. Yet such a restriction may not be followed easily by residents whose concept of spray dryness may vary. Meanwhile, any extended REI may not be practical for residents who are not willing to stay out of their own treated property. An added concern here is that young children could be exposed to surface residues through the hand-to-mouth route, for which the exposure is often considered to be major and impractical to mitigate, if not difficult to assess. For residents or a community at large, sometimes another exposure scenario of concern is when they are exposed to the water residues in private or public pools treated with an antimicrobial. One mitigation option is to reduce the application rate, which like for many other applications may compromise the treatment efficacy. Another option is, much like for many applications made in school buildings, to treat the pools at a certain time of the day or on a certain day of the week with sufficient lapse until the next reentry. This and the preceding examples are but a very few representing the many postapplication expos­ ure scenarios for which EMMs are seemingly less feasible or enforceable, compared to those for workers. Note that in many instances, the user exposure from the pesticide appli­ cation on a residential property is considered to be less of a concern, or even negligible, when compared to the post­ application exposure involved, in that the latter scenario represents a less controllable or less preventable setting. One mitigation option considered to be more enforce­ able is the one specifically intended for bystander workers and for those residents located sufficiently close to a field treated (or being treated) with a fumigant (or an otherwise highly volatile pesticide). This is the use of buffer zones, which typically are based on predictions from some air dispersion models. Insofar as the air dispersion model used has been qualified for assessing the inhalation exposure involved, its use for calculating the buffer zones should be considered acceptable. Naturally, along with the buf­ fer zones comes their posting requirements. Posting may

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be useful for mitigating exposure to nonvolatile pesti­ cides as well, such as when these chemicals are applied to structures or places where residents and bystanders may frequent.

54.4  Cases of pesticide EMMs actually used Not many actual cases of pesticide EMMs are thoroughly documented by regulatory agencies. One reason for such deficiency or inefficiency is that in practice, any exposure mitigation attempt following a pesticide risk assessment is a negotiation process at best. The reality is that only a very few of these attempts need to be (well) documented. By the time the proposed or revised EMMs are accepted through negotiation among regulatory agencies, registrants, and interested parties, they become part of the new language on the revised product label(s). Any mitigation attempts or proposals made in between are usually not considered offi­ cial or formal and hence are neither well documented nor immediately available to individuals who are not among the negotiating parties. In fact, a mitigation option that registrants have, and frequently exercise, is to petition for an opportunity to conduct and submit additional studies to show, for instance, a chemical-specific dermal absorption rate lower than the high default rate that has been used in the pesticide exposure assessment. Naturally, if the dermal absorption rate from an animal study is later determined to be considerably lower than the default value used, the selection of EMMs will normally be affected.

54.4.1  EMMs Actually Used by the U.S. EPA A recent case of EMMs actually used by the U.S. EPA is the federal agency’s new requirement of safety measures for all soil fumigant pesticides as a class. The U.S. EPA (2008b) now requires that for new fumigant labels appear­ ing on the market around 2010, in some situations fumi­ gant users will need to post buffer zones to help ensure that bystanders can stay out of areas that are determined to be vulnerable to excessive inhalation exposure. The federal agency’s REDs are based on a series of intensive pesticide risk assessments. Their REDs for the fumi­ gants chloropicrin, dazomet, metam sodium/potassium, and methyl bromide all included a suite of mitigation measures designed to work collectively to reduce expo­ sures, enhance safety, and facilitate compliance as well as enforcement. Included in this suite of measures are buffer zones, posting requirements, fumigant management plans, good agricultural practices, worker protections, steward­ ship/training programs, and emergency preparedness/ response measures. Another recent case by the U.S. EPA (2008c) is their newly proposed mitigation measures intended to help

Hayes’ Handbook of Pesticide Toxicology

reduce human exposures from use of the 10 rodenticides: brodifacoum, bromadiolone, bromethalin, chlorophaci­ none, cholecalciferol, difenacoum, difethialone, diphaci­ none, warfarin, and zinc phosphide. This proposal, which represents the U.S. EPA’s major effort in minimizing chil­ dren’s exposure to rodenticide products used on residential properties, was issued following a comprehensive assess­ ment of the human health and ecological effects and the benefits involved. In that rodenticide mitigation proposal, the U.S. EPA requires that all those products to be available over the counter be sold only in bait stations meeting the fed­ eral agency’s new requirements. In addition, the federal agency requires that the packaging, distribution, and sales be further restricted for the four second-generation anti­ coagulants brodifacoum, bromadiolone, difenacoum, and difethialone, so that the purchases of these products can be prevented on the consumer market. The four rodenticides have been concluded by the U.S. EPA to pose the greatest risk to wildlife. For pesticides of other classes, which are largely non­ volatile in nature (as for the reason stated in Section 54.3.2), the U.S. EPA’s practice has been to include, where appropriate, EMMs in their REDs, IREDs (i.e., interim REDs), or other related pesticide risk assessment docu­ ments (e.g., U.S. EPA, 1998b, 2002, 2004a,b, 2005b). It has been the federal agency’s practice to consider various levels of worker protection for the purpose of increasing the (calculated) MOEs to the target levels, where deemed appropriate or necessary. The U.S. EPA typically will assess all worker exposures with baseline personal protection first, and then add additional PPE or other types of EMMs using a tiered approach to obtain an appropriate MOE. For example, the exposures and hence the MOEs may be calcu­ lated for mixer/loaders handling a pesticide both with and without gloves worn, even though the labels do not specify the use of such PPE for handling the pesticide under evalu­ ation. Note that for the case just given, registrants could counterpropose to double the estimated MOE to the bench­ mark by using another option that they consider to be more practical, such as by lowering the original maximum label rate by half together with the assurance that treatment effi­ cacy would not be significantly compromised.

54.4.2  EMMs Actually Used in California At the state level in California, cases of EMMs actually used by CDPR are fewer than those adopted by the U.S. EPA. The reason for this is perhaps threefold: The state department not only follows closely the product labels registered with the U.S. EPA and focuses more on use sce­ narios unique to the California setting, but also can work relatively more closely with registrants to resolve many registration and mitigation issues beforehand, as the focus here is primarily on uses within a single state.

Chapter | 54  Mitigation Measures for Exposure to Pesticides

One rather unique California case is that in which two different REIs were adopted by CDPR for reentry to grape vineyards treated with methomyl in different months of the same year (Dong et al., 1992). The use of the two different REIs came about because foliar residue data collected by CDPR indicated a considerably shorter half-life (2 days) for methomyl applied to the California grapes in May through July than in later months of the same year (4 days). Another example for use of REI as an actual EMM in California is that involving the use of DEF (tribufos). In this DEF case (Formoli, 2002), the REI imposed on harvest activities was not finalized until after incorporation of the dermal absorption data from a primate study, which was submitted by registrants many months after CDPR completed the risk characterization document. Mitigation options of other types have also been used by CDPR. These are measures usually proposed during some (late) stage of the pesticide risk assessment process. For example, the cycloate mitigation proposal presented to the registrants by CDPR (Meierhenry, 1995a) in 1995 included the use of either an acceptable closed system or full-body clothing for mixer/loaders. The full-body cloth­ ing would require an approved full-face respirator dur­ ing mixing/loading, in addition to wearing normal work clothes (i.e., long pants, a long-sleeved shirt, and shoes plus socks) and chemical-resistant gloves. As supported by the DPF data presented in Table 54.1, these types of miti­ gation measures would reduce the total exposure at issue at least 10-fold. As another case of non-REI type EMMs used by CDPR, in 1999 registrants were sent notices of cancella­ tion for several DDVP (dichlorvos) pest strip products used in California. The action was taken after the CDPR risk assessment found inadequate MOEs for children in residences and other indoor areas where these pest strips might be used. The notices were issued with the under­ standing that cancellation of most of the pest strip products could be avoided if registrants removed certain uses from the product labels (CDPR, 1999). Another example of this non-REI type occurred in 2003, when California County Agricultural Commissioners were all given notices of proposed mitigation options for dormant sprays of pesticides, with a specific aim at reduc­ ing offsite movement of residues from treated orchards to improve water quality (Sanders, 2003). These mitigation options included, but were not limited to: maintenance of a vegetative buffer strip to ensure a minimum of 10 feet width from field edge to sensitive aquatic sites; no applica­ tion within 100 feet upslope of any sensitive aquatic site; and reduction in maximum application rate or in number of applications. Similar mitigation proposals leading to the reduction of runoff quantity seem to have prevailed to this date, including the more recent one for the pyrethroid insecticide lambda-cyhalothrin (He et al., 2008).

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Last but not least, there are two cases particularly unique to handler exposures in California. These are the EMMs used for handler exposures to EPTC and methyl bromide, where in addition to PPE and engineering con­ trols, other options were used in order to adequately comply with the mitigation requirements. Due to their uniqueness, but more for illustration purposes, these two cases are presented in greater detail in the following, final two sections of this chapter.

54.4.2.1  The EPTC Case The CDPR risk assessment (Meierhenry, 1995b) estab­ lished the following NOELs for EPTC: for acute expos­ ure, 20 mg per kg of body weight (BW) per day based on neurotoxicity in rats; for subchronic (seasonal), 700 g/kg  BW/day based on nasal cavity degeneration/hyperplasia in rats; and for chronic (annual), 500 g/kg BW/day based on neuromuscular degeneration in rats. With these NOELs, the MOEs were calculated to be from 90 to 6974 for acute and chronic exposures, which were considered to be acceptable as they were all either close to or exceeding 100. However, some of the MOEs calculated for subchronic exposures were as low as 17, depending on the work task or product formulation involved. Accordingly, CDPR staff (Thongsinthusak, 1998) pro­ posed several possible EMMs to reduce the seasonal der­ mal and inhalation exposures for EPTC handlers. These EMMs included the use of PPE, a respirator, and limited work hours for a certain exposure period. During the com­ ment period, EPTC registrants suggested that the amount of EPTC handled be used in lieu of the work hours allowed, arguing that the amount of EPTC that they specified for maximum handling by workers was equivalent in mitiga­ tion efficacy to the work hours per time period proposed by CDPR. Table 54.3 summarizes the EMMs that were final­ ized for seasonal exposures to the various work tasks and formulations under consideration. In 1995, these measures were incorporated into federal product labels as additional requirements specifically for California handlers.

54.4.2.2  The Methyl Bromide Case The CDPR risk assessments (Lim, 2002, 2003) estab­ lished the NOELs for methyl bromide as follows: for acute ex­posure, 40 parts per million (ppm) based on develop­ mental toxicity in pregnant rabbits; for subchronic, 5 ppm based on neurotoxicity in dogs; and for chronic, 0.3 ppm (estimated NOEL) based on nasal epithelial hyperplasia/ degeneration in rats. The reference concentrations (i.e., human equivalent NOELs divided by safety/uncertainty factors assumed) used as benchmarks were 210 parts per billion (ppb) for acute exposure (adult and children); 16 ppb (adult) and 9 ppb (children) for subchronic; and 2 ppb (adult) and 1 ppb (child) for chronic (Lim, 2002, 2003).

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Table 54.3  Mitigation Measures for Seasonal (Subchronic) Exposure of EPTC Handlers in Californiaa,b Work task/formulation

Mitigation measure(s)

Handlers

Coveralls plus a half-face respiratorc

Handlers using center   pivot (CP)

Chemical-resistant full-body protective clothing and a halfface respiratorc

Liquids   Mixer/loader

75 gal handled/day, or 500 gal/21-day period

  Applicator using CP

20 gal handled/day, or 40 gal/21day period

 Applicator in an   enclosed cab

40 gal handled/day, or 280 gal/21-day period

 Applicator using other  type of application equipment

30 gal handled/day, or 210 gal/21-day period

Granules   Handlers

100 lb handled/day, or 1000 lb/21-day period

a

Adapted from Thongsinthusak and Frank (2007), courtesy of ACS (American Chemical Society Symposium Series 951); EPTC  s-ethyl dipropylthiocarbamate. b For product labels indicating “The operator of the property shall include in their pesticide use records the name(s) of person(s) that handled the product for each application.” c The respirator used must be equipped with one of the following: (1) an organic vapor (OV) removing cartridge with a prefilter approved for pesticides (MSHA/NIOSH approval no. prefix TC-23C); (2) a canister approved for pesticides (MSHA/NIOSH approval no. prefix TC-14G); (3) a NIOSH-approved respirator with an OV cartridge; or (4) a canister with any N, R, P, or HE prefilter.

Accordingly, several possible EMMs were proposed by CDPR staff for the acute (Gibbons and Thongsinthusak, 2000) and seasonal exposures (Thongsinthusak and Frank, 2003). These EMMs are summarized in Table 54.4. Presently, all of the proposed EMMs have been incorp­ orated as state regulations (CCR, 2008b). Note that the exposure mitigation for chronic effects of methyl bro­ mide was not addressed, as in other similar cases, mainly because there was a general understanding that if and when the subchronic exposures were sufficiently reduced, so would the related chronic exposures. Table 54.4 shows the maximum daily work hours allowed per application method and rate for both the acute and subchronic exposures. In essence, for acute exposure, methyl bromide handlers are allowed to work from 2 to 7 h per workday without the use of a respir­ator. For subchronic exposure, however, California work­ ers may handle the fumigant according to the speci­ fied work hours but must wear those NIOSH-certified re­spiratory devices recommended by manufacturers for use specifically in an atmosphere having less than 5 ppm methyl bromide. The additional mitigation requirements specified in the state regulations (CCR, 2008b), but not included in Table 54.4, are as follows: specific field fumigation methods, field fumigation notification, buffer zones, and techniques for tarpaulin cutting, removal, and repair. What seems to be particularly unique about the methyl bromide regulations is that they provide flexibility on cal­ culation of maximum work hours allowed and on adjust­ ment of work hours that are based on air concentrations and actual application rates. The required calculation pro­ cedures are as follows:

Table 54.4  Maximum Work Hours Allowed for Acute and Seasonal (Subchronic) Exposures to the Methyl Bromide Active Ingredient (AI)a Application method and work taskb

Maximum daily work hoursc

Maximum lb AI/acre

Acute

Seasonal

4d

9d

4d

9d

4d

10d

7d

No limitatione

A. Nontarpaulin/shallow/bed   Tractor equipment driving

200

  Supervising B. Nontarpaulin/deep/broadcast   Tractor equipment driving   Supervising

400

Chapter | 54  Mitigation Measures for Exposure to Pesticides

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Table 54.4  (Continued) Application method and work taskb

Maximum daily work hoursc

Maximum lb AI/acre

Acute

Seasonal

4d

8d

3d

4d

3d

4d

C. Tarpaulin/shallow/broadcast   Tractor equipment driving   Shoveling, copiloting

400

  Supervising   Tarpaulin cutting

e

No limitatione

7

e

No limitationf

4d

No limitation

4

  Tarpaulin removal D. Tarpaulin/shallow/bed   Tractor equipment driving   Shoveling, copiloting

250

  Supervising

d

8d

4

d

8d

4d

8d

3d

4d

3

d

4d

2d

5d

d

5d

4

E.  Tarpaulin/deep/broadcast   Tractor equipment driving   Shoveling, co-piloting

400

  Supervising F.  Drip system: hot gas   Applicator   Supervising

225

2

a

Adapted from Thongsinthusak and Frank (2007), courtesy of ACS (American Chemical Society Symposium Series 951). Specific requirements are described in the state regulations (CCR, 2008b). c The maximum hours allowed are in a maximum three workdays per calendar month for acute exposure, but are based on a 24-h period for seasonal; also, handlers are required to wear a half-face respirator for seasonal exposure but not for acute. d Work hours can be adjusted by using the formula given in Eq. (6) in the text. e Same work hours for these two handler tasks are used for application methods D, E, and F; the following exception is given for seasonal exposure: An employee may perform this activity without a half-face respirator provided that (s)he does not work more than 1 h in a 24-h period, and the maximum 1-h work limitation may be increased in accordance with the formula given in Eq. (6) in the text. f Same work hours are applied for tarpaulin removal for application methods D, E, and F; the following exception is given for seasonal exposure: An employee may perform this activity without a half-face respirator provided that (s)he does not work more than 3 h in a 24-h period, and the maximum 3-h work limitation may be increased in accordance with the formula given in Eq. (6) in the text. b

Conclusion

Determination of daily work hours for acute exposures:



(Maximum work hours)  [(210 ppb  24 h )  (95th percentile breathing zone methyl bromide concentration in ppb, not the time­weighted average)]



(4)

Determination of daily work hours for seasonal exposure:



(Maximum work hours)  [(16 ppb  24 h )  (arithmetic mean breathing zone methyl bromide concentration in ppb, not thee time­weighted average)]



(5)

Revision of maximum daily work hours permitted by the regulations:



(Maximum work hours)  [(maximum application rate for method used )  (maximum work hours in a 24­h exposure period) ÷ (actual application rate used )]

(6)

Exposure mitigation, particularly for pesticides, is an inte­ gral part of health risk prevention. Numerous effective exposure mitigation measures (EMMs) have been devel­ oped or advocated by many health professionals involved in pesticide risk assessment. Accordingly, this chapter was written to describe the evolution and use of the EMMs involved. It first identifies five elements fundamental to the development of pesticide EMMs, including evaluation for potential adverse health effects and the use of EMMs at a reasonable cost. The chapter then provides several criteria for more practical EMMs used for the protection of three groups comprising most of the exposed population: pesti­ cide handlers/users, fieldworkers, and residents/bystanders. For handlers/users, the most effective options continue to be PPE (personal protective equipment) and engineering controls. With fieldworkers, extended REI (restricted entry interval) appears to be the most practical, if not the most effective, EMM. For residential exposure, especially when

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children are involved, one effective EMM is to reformulate the pesticide or to redesign the product packaging so that the user or nonuser would receive less exposure when or after the product is applied in a residential area. The chap­ ter concludes with several real cases of pesticide EMMs used in the United States, including those used specifically in the state of California. The two cases particularly unique to handler exposures in California are the EMMs used for handler exposures to methyl bromide and EPTC (s-ethyl dipropylthiocarbamate). For these two pesticides, other measures along with PPE and engineering controls were required in order to adequately comply with the mitigation requirements.

Acknowledgments The opinions expressed in this chapter represent those of the authors and do not necessarily reflect the views or poli­ cies of the California Department of Pesticide Regulation or of the California Environmental Protection Agency; and the mention of commercial products does not constitute an endorsement or recommendation for their use.

References Andrews, C. M. (2000). “Worker Health and Safety Branch Policy on the Statistical Analysis for Dislodgeable Foliar Residue Data. HSM-00011,” Worker Health and Safety Branch, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California. ANSI (American National Standards Institute) (2006). “American National Standard for Respiratory Protection. Respirator Use – Physical Qualifications for Personnel Use,” ANSI, American Industrial Hygiene Association, Fairfax, Virginia. Aprea, C., Sciarra, G., Sartorelli, P., Desideri, E., Amati, R., and Sartorelli, E. (1994). Biological monitoring of exposure to organophosphorus insecticides by assay of urinary alkylphosphates: Influence of protec­ tive measures during manual operations with treated plants. Int. Arch. Occup. Environ. Health 66, 333–338. ASAE (American Society of Agricultural Engineers) (1998a). “Agricultural Cabs – Environmental Air Quality. Part 1: Definitions, Test Methods, and Safety Practices. ASAE S525-1.1,” ASAE, St. Joseph, Michigan. ASAE (American Society of Agricultural Engineers) (1998b). “Agricultural Cabs – Environmental Air Quality. Part 2: Pesticide Vapor Filters – Test Procedure and Performance Criteria. ASAE S525-2,” ASAE, St. Joseph, Michigan. Brouwer, D. H., de Haan, M., and van Hemmen, J. J. (2000). Modeling re-entry exposure estimates: Techniques and application rates. In “Worker Exposure to Agrochemicals – Methods for Monitoring and Assessment” (R. C. Honeycutt and E. W. Day Jr., eds.), pp. 119–138. Lewis Publisher, Boca Raton, Florida. CCR (California Code of Regulations) (2008a). “Personal Protective Equipment. Title 3 (Food and Agriculture), Section 6738,” California Office of Administrative Law, Sacramento, California.

Hayes’ Handbook of Pesticide Toxicology

CCR (California Code of Regulations) (2008b). “Field Fumigation. Title 3 (Food and Agriculture), Section 6784,” California Office of Administrative Law, Sacramento, California. CDPR (California Department of Pesticide Regulation). (1999). News Release (February 19): “DPR Moves to Cancel Registration of Pest Strips,” Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California. http:// www.cdpr.ca.gov/docs/pressrls/archive/1999/990218.htm. CFAC (California Food and Agricultural Code)—BDPA (The Birth Defect Prevention Act). (1984). Sections 13121–13135. California Codes, Official California Legislative Information, Legislative Counsel of California, Sacramento, California. http://www.leginfo.ca.gov/ cgi-bin/displaycode?sectionfac&group13001-14000&file 13121-13135. CFR (Code of Federal Regulations) (2008). “Approval of Respiratory Protective Device. Title 42, Part 84,” Office of the Federal Register, National Archives and Records Administration, U.S. Government Printing Office, Washington DC. del Valle, M. (1999). “Protocol for the Handling of Dislodgeable Foliar Residue (DFR) Samples,” Worker Health and Safety Laboratory, California Department of Food and Agriculture, Sacramento, California. Dong, M. H. (1990). “Dermal Transfer Factor for Cotton Scouts. HSM-90001,” Worker Health and Safety Branch, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California. Dong, M. H. (2004). Letter to Editor: Regul. Toxicol. Pharmacol. 39, 239. Dong, M. H., Krieger, R. I., and Ross, J. H. (1992). Calculated reentry interval for table grape harvesters working in California vine­ yards treated with methomyl. Bull. Environ. Contam. Toxicol. 49, 708–714. Durham, W. F., and Wolfe, H. R. (1962). Measurement of the exposure of workers. Bull. Wld. Hlth. Org. 26, 75–91. Edmiston, S., O’Connell, L., Bissell, S., and Conrad, D. (2002). “Guidance for Determination of Dislodgeable Foliar Residue. HS-1600,” Worker Health and Safety Branch, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California. Faustman, E. M., and Omenn, G. S. (2001). Risk assessment. In “Casarett and Doull’s Toxicology: The Basic Science of Poisons” (C. D. Klaassen, ed.) Sixth Edition, pp. 83–104. McGraw-Hill, New York, New York. Fenner-Crisp, P. A. (2001). Risk assessment and risk management – The regulatory process. In “Handbook of Pesticide Toxicology” (R. I. Krieger, ed.) Second Edition Vol. I: Principles, pp. 681–689. Academic Press, San Diego, California. Formoli, A. (2002). “Tribufos Mitigation. HSM-02003,” Worker Health and Safety Branch, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California. Gerritsen-Ebben, R. (M. G.), Brouwer, D. H., and van Hemmen, J. J. (2007). “Effective Personal Protective Equipment (PPE): Default Setting of PPE for Registration Purposes of Agrochemical and Biocidal Pesticides,” TNO Report V7333. TNO Quality of Life, TNO Chemistry, Food & Chemical Risk Analysis, Chemical Exposure Assessment, A J Zeist, the Netherlands. Gibbons, D., and Thongsinthusak, T. (2000). “Recommended Worker Safety Mitigation Measures for Methyl Bromide Soil Fumigation Regulations. HSM-00014,” Worker Health and Safety Branch, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California.

Chapter | 54  Mitigation Measures for Exposure to Pesticides

Goldstein, B. D. (2003). Risk characterization and the red book. J. Hum. Ecolog. Risk Assess. 9, 1283–1289. Gosselin, P. G. (2001). “Department Policy on the Risk Management and Mitigation Process. HSM-01005,” Executive Office, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California. Grieshop, J. I. (1988). Protective clothing and equipment – Beliefs and behavior of pesticide users in Ecuador ASTM STP 989. In “Performance of Protective Clothing: Second Symposium” (S. Z. Mansdorf, R. Sager, and A. P. Nielsen, eds.), pp. 802–809. American Society for Testing and Materials (ASTM), Philadelphia, Pennsylvania. He, I.-M., Troiano, J., Wang, A., and Goh, K. (2008). Environmental chemistry, ecotoxicity, and fate of lambda-cyhalothrin. Rev. Environ. Contam. Toxicol. 195, 71–91. Iwata, Y., Knaak, J. B., Spear, R. C., and Foster, R. J. (1977). Worker reentry into pesticide-treated crops. I. Procedure for the determination of dislodgeable pesticide residues on foliage. Bull. Environ. Contam. Toxicol. 18, 649–655. Krebs, B., Maasfeld, W., Schrader, J., Wolf, R., Hoernicke, E., Nolting, H. G., Backhaus, G. F., and Westphal, D. (2000). Uniform principles for safeguarding the health of workers re-entering crop growing areas after application of plant-protection products. In “Worker Exposure to Agrochemicals – Methods for Monitoring and Assessment” (R. C. Honeycutt and E. W. Day Jr., eds.), pp. 107–118. Lewis Publishers, Boca Raton, Florida. Krieger, R. I., Ross, J. H., and Thongsinthusak, T. (1992). Assessing human exposure to pesticides. Rev. Environ. Contam. Toxicol. 128, 1–15. Liebman, A. K., Juarez, P. M., Leyva, C., and Corona, A. (2007). A pilot program using promotoras de salud to educate farmworker families. J. Agromed. 12, 33–43. Lim, L. O. (2002). “Risk Characterization Document for Methyl Bromide: Inhalation Exposure (Volume I),” Technical Report Number RCD-2002-03. Medical Toxicology Branch, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California. Lim, L. O. (2003). “Risk Characterization Document for Methyl Bromide: Inhalation Exposure (Addendum to Volume I),” Medical Toxicology Branch, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California. Meierhenry, E. F. (1995a). “Risk Characterization Document for Cycloate (Ro-Neet),” Medical Toxicology Branch, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California. Meierhenry, E. F. (1995b). “Risk Characterization Document for EPTC (s-Ethyl Dipropylthiocarbamate),” Medical Toxicology Branch, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California. NRC (U.S. National Research Council) (1983). “Risk Assessment in the Federal Government: Managing the Process. Committee on the Institutional Means for Assessment of Risks to Public Health, Commission on Life Sciences, NRC,” National Academy Press, Washington DC. Palis, F. G., Flor, R. J., Warburton, H., and Mahabub, H. (2006). Our farmers at risk – Behavior and belief system in pesticide safety. J. Public Health 28, 43–48. Popendorf, W. J. (1985). Advances in the unified model for re-entry haz­ ards in dermal exposure related to pesticide in use. In “Discussion of Risk Assessment” (R. C. Honeycutt, G. Zweig, and N. N. Ragsdale,

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eds.), ACS Symposium Series 273, pp. 323–340. American Chemical Society (ACS), Washington DC. Sanders, J. S. (2003). “Proposed Mitigation Options for Dormant Sprays. EM-0302,” Environmental Monitoring Branch, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California. Thongsinthusak, T. (1998). “EPTC Exposure Mitigation Document. HSM-98002,” Worker Health and Safety Branch, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California. Thongsinthusak, T., Brodberg, R. K., Ross, J. H., Gibbons, D., and Krieger, R. I. (1991). “Reduction of Pesticide Exposure by Using Protective Clothing and Enclosed Cabs. HS-1616,” Worker Health and Safety Branch, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California. Thongsinthusak, T., and Frank, J. P. (2003). “Mitigation Measures for Seasonal Exposures of Agricultural Workers to Methyl Bromide During Soil Fumigations. HSM-03012,” Worker Health and Safety Branch, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California. Thongsinthusak, T., and Frank, J. P. (2007). Developing pesticide exposure mitigation strategies. In “Assessing Exposures and Reducing Risks to People from the Use of Pesticides” (R. I. Krieger, N. Ragsdale, and J. N. Seiber, eds.), ACS Symposium Series 951, pp. 98–110. American Chemical Society (ACS), Washington DC. Thongsinthusak, T., and Ross, J. (1994). “Protection Provided by the Closed System. HSM-94006,” Worker Health and Safety Branch, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California. Thongsinthusak, T., Ross, J. H., and Meinders, D. (1993). “Guidance for the Preparation of Human Pesticide Exposure Assessment Documents. HS-1612,” Worker Health and Safety Branch, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, California. U.S. EPA (U.S. Environmental Protection Agency). (1995). Guidance for Risk Characterization. Science Policy Council, U.S. EPA, Washington DC. http://www.epa.gov/osa/spc/pdfs/rcguide.pdf. U.S. EPA (U.S. Environmental Protection Agency). (1998a). Notice to Manufacturers, Producers, Formulators and Registrants of Agricultural Pesticides. Pesticide Registration (PR) Notice 98-9. Office of Prevention, Pesticides and Toxic Substances, U.S. EPA, Washington DC. U.S. EPA (U.S. Environmental Protection Agency). (1998b). Reregistration Eligibility Decision (RED) for Methomyl. Case 0028. EPA 738-R-98-021. Office of Prevention, Pesticides and Toxic Substances, U.S. EPA, Washington DC. U.S. EPA (U.S. Environmental Protection Agency). (1999a). Reregistration Eligibility Decision (RED) for EPTC (s-Ethyl Dipropylthiocarbamate). Case 0064. EPA 738-R-99-006. Office of Prevention, Pesticides and Toxic Substances, U.S. EPA, Washington DC. U.S. EPA (U.S. Environmental Protection Agency). (1999b). Reregistration Eligibility Decision (RED) for Chlorothalonil. Case 0097. EPA 738-R-99-004. Office of Prevention, Pesticides and Toxic Substances, U.S. EPA, Washington DC. U.S. EPA (U.S. Environmental Protection Agency). (2000). Policy Number 3.1: Agricultural Transfer Coefficients (revised August 7). Science Advisory Council for Exposure, Health Effects Division, Office of Pesticide Programs, U.S. EPA, Washington DC.

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U.S. EPA (U.S. Environmental Protection Agency). (2002). Interim Reregistration Eligibility Decision (IRED) for Naled. Case 0092. EPA 738-R-02-008. Office of Prevention, Pesticides and Toxic Substances, U.S. EPA, Washington DC. U.S. EPA (U.S. Environmental Protection Agency). (2004a). Reregistration Eligibility Decision (RED) for Cycloate. Case 2125. EPA 738-R-04-014. Office of Prevention, Pesticides and Toxic Substances, U.S. EPA, Washington DC. U.S. EPA (U.S. Environmental Protection Agency). (2004b). Interim Reregistration Eligibility Decision (IRED) for Diazinon. Case 0238. EPA 738-R-04-006. Office of Prevention, Pesticides and Toxic Substances, U.S. EPA, Washington DC. U.S. EPA (U.S. Environmental Protection Agency) (2005a). “How to Comply with the Worker Protection Standard for Agricultural Pesticides: What Employers Need to Know,” Office of Pesticide Programs, U.S. EPA, Washington DC. U.S. EPA (U.S. Environmental Protection Agency). (2005b). Reregistration Eligibility Decision (RED) for 2,4-D. Case 0073. EPA 738-R-05-002. Office of Prevention, Pesticides and Toxic Substances, U.S. EPA, Washington DC. U.S. EPA (U.S. Environmental Protection Agency). (2008a). Reregistration Eligibility Decision (RED) for Methyl Bromide (Soil and Nonfood Structural Uses). Case 0335. EPA 738-R-08-005.

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Office of Prevention, Pesticides and Toxic Substances, U.S. EPA, Washington DC. U.S. EPA (U.S. Environmental Protection Agency). (2008b). Implementation of Risk Mitigation Measures for Soil Fumigant Pesticides. Office of Pesticide Programs, U.S. EPA, Washington DC; http://www.epa.gov/oppsrrd1/reregistration/soil_fumigants/. U.S. EPA (U.S. Environmental Protection Agency). (2008c). Final Risk Mitigation Decision for Ten Rodenticides. Office of Pesticide Programs, U.S. EPA, Washington DC; http://www.epa.gov/oppsrrd1/ reregistration/rodenticides/finalriskdecision.htm. van Hemmen, J. J., Brouwer, D. H., and de Cock, J. S. (2001). Greenhouse and mushroom house exposure. In “Handbook of Pesticide Toxicology” (R. I. Krieger, ed.) Second Edition (Volume I: Principles), pp. 457–478. Academic Press, San Diego, California. Whitmyre, G. K., Ross, J. H., Lunchick, C., Volger, B., and Singer, S. (2003). Biphasic dissipation kinetics for dislodgeable foliar residues in estimating postapplication occupational exposures to endosulfan. Arch. Environ. Contam. Toxicol. 46, 17–23. Willis, G. H. and McDowell, L. L. (1987). Pesticide persistence on foli­ age. Rev. Environ. Contam. Toxicol. 100, 23–73. Wolfe, H. R. (1976). Field exposure to airborne pesticides. In “Air Pollution from Pesticides and Agricultural Processes” (R. E. Lee Jr. ed.). CRC Press, Cleveland, Ohio.

Chapter 55

Communicating Safe Pesticide Use Allan S. Felsot Washington State University, Richland, Washington

Consider the following anecdote as an illustration of the challenges faced by crop and health protection special­ ists (a term I intend to include all persons, agencies, and industries involved in all aspects of pest management). In my role as extension specialist for environmental toxicol­ ogy, I was requested by a group representing state noxious weed control boards to make a presentation countering the seemingly proliferating claims of harm from use of herbicides for controlling invasive and noxious weeds on natural lands as well as along roadsides. In particular, the organizers wanted to know how they could more effec­ tively communicate their story. Undoubtedly, they were motivated by both their mission to halt further degradation of lands by invasive plants as well as their confidence in the benefit of synthetic chemical pesticides to achieving their goal. The organizers wished to headline my presen­ tation with this somewhat unwieldy title: “The Semantics of Risk Management: Winning the Communication Battle for Sensible Pesticide Use.” Perhaps faster than a speed­ ing bullet, an iconic Pogo cartoon aphorism came to mind: “We have met the enemy, and he is us.” Clearly, the pest management technology enterprise needs to communicate more effectively what it is doing and engage the public honestly about technological risks. Such need is strongly suggested by the hundreds of articles writ­ ten each year about risk communication, most likely with an objective of improving the process. An aggregate of billions of dollars are spent annually to identify pesticide hazards, assess associated risks, and develop reasonable regulations of the commercialized products. An objective observer at the least should conclude that the methodol­ ogy for ferreting out hazards, quantitating risks, and imple­ menting regulations has improved greatly from 40 years ago. Nevertheless, public concerns about the technology, which has expanded to include pest protective characters of agricultural crops subjected to transgenic engineering methods, have not tapered but arguably grown. So, all the good money spent in developing comparatively safer com­ pounds is not enhancing communication that effectively Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

consoles the public about negligible risks, yet alone the shared benefits from such technologies. My objective herein is to make a case that risk commun­ ication is needed by the direct users of pesticide technol­ ogy (not just the manufacturers or governmental regulators) and to improve the effectiveness of such communication. Based on personal experiences interlaced with information from the risk communication literature, I will explore who the real audience may be for risk communication mes­ sages from pesticide users, how audiences and constitu­ encies perceive risk, and what we know about consumer perceptions of chemical technology risks. I will challenge the reader to question what they know about the technol­ ogy they use and whether they are ready to provide accu­ rate information. Finally, I will suggest ways to improve communication with a user’s constituents. In connection with the latter need, I will address how precautionary prin­ ciple (PP) advocacy may undermine efforts to effectively communicate safe pesticide use, but also suggest a strat­ egy for incorporating some of the PP precepts into users’ messages.

55.1  Why pesticide technology users need a risk communication strategy The phrase “wealthier, healthier, and more worried” aptly describes the typical Western consumer today (Slovic, 1991). Ironically, as newspaper headlines have reported Centers for Disease Control reports of increased longevity with overall improvements in health, people seemingly are more stressed out and concerned about technology in soci­ ety. An explanation of why good news might be ignored has been succinctly listed as follows (Slovic, 1991): Increased technical capability of measuring trace lev­ els of chemicals (and thus finding synthetics just about everywhere, including in our bodies)

l

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Increased dependence on complex, powerful technol­ ogies (e.g., large region electrical blackouts) l Occurrence of catastrophic mishaps (nuclear plant core “meltdowns” such as Chernobyl in the former Soviet Union) l Increased litigation (tracked by the rise in toxic tort lawsuits) l More to lose of one’s personal wealth or station pro­ motes tendency to be more cautious l Living with complex, interconnected systems underlain by a nearly incomprehensible technology l A 24/7 media exposure that focuses on risks l Benefits of technology taken for granted (e.g., the expectation that water and electricity will always flow from the tap or switch) l Frustration over lack of control and the involuntary nature of many risks l

Any one of the reasons for general uncertainty with regard to modern living in general certainly is applicable to pes­ ticide technology. A picture is worth a thousand words, so a brief perusal of newspaper headlines makes evi­ dently clear that pesticide technology users have a lot of explaining to do. Below is a selected sampling of verba­ tim newspaper headlines that concisely emphasize hazard and by implication high risk from use of pesticides. After many years of collecting newspaper and news magazine headlines, I’ve grouped them into examples of fear, con­ fusion, mistrust, and hope. Not all necessarily relate to pesticide technology directly, but rather convey a message in general about validity or confusion over hazards. For example, headlines such as “Inactivity reportedly kill 1.9 million a year” seem counterintuitive to conventional wis­ dom and are therefore offered as an example of confusion (Table 55.1). From this panoply of information, consumers who are likely to interact first hand with urban pest control opera­ tors cannot possibly be expected to have a comfortable cer­ tainty with pesticide technology. Given that less than 2% of the U.S. population has a direct connection with agricul­ tural producers, pest control technology applied to scales larger than their own yards or household may be hard to fathom.

55.2  Don’t blame the media The headline examples show a print and now on-line media that could be characterized as sound bite driven, conflict seeking, and presenting stories without context. To blame the media, however, will not improve risk communica­ tion nor help craft useable messages. Such blame is mis­ informed by conventional wisdom of a negative bias in the media and a mistaken impression that the public wholly depends on the television and print media for information. However, media coverage more likely entrenches preexisting

Hayes’ Handbook of Pesticide Toxicology

prejudices and/or worries (except when a risk is novel). For example, in an outbreak of cryptosporidium in a pub­ lic water supply, survey respondents were more likely to use the media for information if they already had concerns about becoming ill (Griffin et al., 1998). Thus, the concern came before any influence of the media. Indeed, the indi­ viduals are more likely to use mass media risk stories for information about risks but resist messages of how worried to be (Dunwoody, 1991). Television and print media often emphasize qualita­ tive (i.e., human interest) aspects of a story, so balancing a story results in an overemphasis on a minority opinion. Also true, however, is the need to entice readers, so stories with a villain underlay many narratives (Dezenhall, 1999). Perhaps the most intense cases of media scrutiny about a pesticide hazard were the reports during 1989 of alar residues in apples. Public policy in reaction to perceived consumer concerns reacted by recommending suspension of the use of alar as a growth regulator. However, analy­ sis of media conduct regarding balance of stories showed that the industry perspective received prominent coverage. If anything, the media spent more attention on the conflict between stakeholders than on the science behind the risk (Friedman et al., 1996). Media of all kinds are dependent on credible and articulate sources of information. The universal accessi­ bility of the Internet now makes such information gather­ ing faster but ideally more informative. Most importantly, media can be a “friend” if sought out, encouraged, and assisted in becoming involved with risk communication programs. Although media is rapidly moving from print to exclusive digital availability, and television news sources have dispersed from the networks to the 24/7 cable out­ lets, the Internet has completely democratized the ability of any organization to get their message to a community. Thus, careful and thoughtful crafting of risk communica­ tion messages by crop and health protection specialists has more outlets than ever before. However, the town hall pub­ lic meeting will still likely be a place for direct one-to-one communication, so preparation with regard to audience background and their perceptions is beneficial to effective communication.

55.3  Who is the appropriate audience for risk communication? The need for a risk communication strategy is somewhat obvious given concerns about the quality of mass infor­ mation flow today. More importantly, perhaps, is deciding who the actual audience for the message really is. If you are a pest control operator or landscape manager relying on pesticide use, the customer is the audience needing honest unbiased information. But in these cases, the customer has already called you with the expectation that you are going

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Table 55.1  Collection of Verbatim Newspaper Headlines Illustrating Common Themes Likely Reflecting or Affecting Public Perceptions of Risk.a Selected examples of fear

Examples of confusion

Examples of mistrust

Examples of hope

“Study finds increased risk of cancer for those using 2,4-D”

“Inactivity reportedly kills 1.9 million a year”

“EPA pushed to protect fish from pesticides”

“38 pesticides barred near salmon waters”

“Cancer rates high in Hispanic ag workers, study says”

“Longer life means more concerns, AARP says; but people older than 50 are generally healthier”

“Bush to ease pesticide reviews”

“State suspends pesticide licenses”

“Lawn-care chemical tied to cancer; dog study may show humans are at risk”

“Having sons could shorten mothers’ life “River buffers worry farmers” spans, researchers say; daughters nurtured to adulthood may prolong mothers’ life”

“Pesticide blood test rule begins in February”

“Pesticide spray concerns rural residents”

“Study warns of vitamin C dangers”

“Survey shows gap between scientists and the public”

“State sampling fish in Lake Chelan for DDT”

“High levels of pesticide found in state streams”

“Health: Soya-based foods may harm male fertility, say scientists”

“EPA faces suit over farm use of “U.S. cancer death rates once-banned insecticide” are found to be falling”

“Large amounts of dangerous pesticide still used”

“State gives warning on fold remedies”

“Pesticide monitoring program data sparse”

“Researchers find pesticide traces in kids”

“Green tea not cure all after all?”

“Consumers, citizens must demand disclosure on biotech foods”

“Tests show high pesticide exposure in farmworkers”

“Environmental groups sue EPA over gene-altered crops”

“The pollution within; toxic chemicals common in body, but how much is too much?” “Toddler dies from pesticide in home” “Home products contaminating waterways” “Deformed frogs are surfacing in Oregon ponds; no human health risk suspected” “Wenatchee schools soil contaminated” “Most state schools use harmful pesticides” “DDT levels highest in Lake Chelan” “Gene altered DNA may be polluting corn” “Half of cities’ water tainted by pesticides; study expresses concern for long-term effects” “Cancer the top killer for those under 85” “Study ties environmental pollution to autism rates” “New corn endangers Monarch butterfly” “Unapproved biotech corn found in taco shells” a

Nearly all headlines were clipped from hard copy or on-line news sources after 2000.

“FDA to test for biotech corn allergy”

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to do what is needed to solve their pest problem. Perhaps such a customer already has faith in the power of technol­ ogy to solve the problem. However, the consumer down the street, or in the case of agricultural lands, the com­ munity of farmworker families often in close proximity to a field, likely need a message tailored to their needs, not being vested in the outcome of the pest control operation. Any uncertainty among the bystanders can be pushed for­ ward to the political level through petition of “grievances.” Widespread cases now involve complaints to city and county governments about weed control along roadsides. The communication needs for crafting risk communication message with these latter agencies may require entirely dif­ ferent objectives as well as content. To illustrate what many pesticide users should be con­ cerned about, consider the story behind this headline in an on-line news site: “Two Efforts Seek to Halt Pesticide Use.” The story reports the combined actions of Long Island environmental advocates and allies in the New York State Legislature to propose a state law that would “ban many lawn chemicals that cause cancer in lab animals.” Such a ban is reminiscent of the problems with practical enforcement of the Delaney Amendment to the FFDCA in 1958. In essence, the amendment forbade setting a toler­ ance for any pesticide residue in processed foods if that chemical had tested positive in a rodent carcinogenic­ ity study. The amendment was essentially unenforce­ able because it did not apply to raw food that eventually might become processed food (such as grapes to raisins). Pertinently, the law essentially resulted in setting a zero risk standard for any detectable residue (NRC, 1987), yet the validity of tagging the label of carcinogen on a chemi­ cal has been challenged when solely based on studies of exposure at the maximum tolerated dose (Ames, 1992). Congressional passage of the FQPA in 1996 rescinded the Delaney Amendment and unified the risk of pesticide resi­ due exposure from raw and processed agricultural com­ modities. Nevertheless, environmental policy regarding the use of urban pesticides has been poised to once again set rules on the basis of uncertain interpretations of one type of toxicological test. Lest the pest control operators face the imposition of restrictions that go beyond EPA’s management decision in issuing registration, they might be wise to consider that perhaps their audience is really councilmen and legislators. Once a law is proposed, efforts to convince the public that pesticide use is actually low risk are wasted because the audience has shifted. Some research suggests that legislators may indeed be a primary audience for a coherent effort in risk communi­ cation if the objective of pesticide users is to avoid onerous and unnecessary regulations. Hazard and risk assessments have been increasingly used to set state and local environ­ mental policies that can affect various technologies, includ­ ing pesticide use. How legislators and their staffs perceive

Hayes’ Handbook of Pesticide Toxicology

risk and the risk assessment process probably influences evaluations of competing information from stakeholders and the resulting decisions (Cohen, 1997). Given that leg­ islators are likely to heavily rely on their staff for informa­ tion regarding hazards of various technologies, legislative staffs logically seem prime audiences for risk communica­ tion long before an issue comes to the rule making body’s floor. A survey of the top three kinds of information that legislators and staff prefer as most useful include “an explanation of how regulatory agencies like the EPA use risk assessment,” “an explanation of the strengths and weaknesses of risk assessments,” and “an explanation of the risks of a particular technology or situation” (Cohen, 1997). While the curiosity for more information is hopeful, responses are decidedly divided when questions focus on the current state of knowledge (and feelings) about likely hazards from exposure to chemicals. For example, 47% and 39%, respectively, disagree or agree that positive carcino­ genicity tests from animal studies are likely to cause cancer in humans. On the other hand, 52% of legislators and staff respondents believed that exposure to a toxic chemical was sure to result in adverse health effects, and 42% disagreed. A somewhat pessimistic view (albeit reflecting the 1990s) prevailed on the future chemical risks to be discovered, with 57% thinking that society has only seen the tip of the iceberg, but 34% disagreeing. From an economic perspec­ tive, 76% of legislators and staff disagreed with the propo­ sition that “it can never be too expensive to reduce the risks associated with chemicals” (Cohen, 1997). Pertinently for thinking about risk communication, differences were some­ what differentiated by political party affiliation.

55.4  Perceptions of risk: why people believe and feel what they do The fact that legislators’ and their staff’s responses are somewhat correlated with their political party affiliation suggests personal bias must be a factor in risk perception. However, this one factor is likely overwhelmed by a host of other factors that have been studied as applicable to the public in general. How one perceives risk is not simply dependent on receiving, processing, and analyzing factual information about an activity. Difficulty in understanding the science behind chemical technology is only a small part of judging a technology as too risky. Thus crafting an effective message requires an understanding of the factors that influence how risk is perceived. In general, research has shown that risks have qualitative factors or dimensions suggesting that people perceive or even feel a risk. Probably the most cited, comprehensive analysis of how people perceive risk was conducted by Slovic (1987) who placed various activities or technologies (“hazards”) along a two-factor risk continuum (unknown/known and

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1177

dread/not dread) that was displayed as a simple x-y graphic (Figure 55.1). The two-factor representation of risk percep­ tions was built on earlier work, and in both studies pesti­ cides were included as an individual hazard (Fischhoff et al., 1978). Each of the factors (i.e., dread risk or unknown risk) was defined at its extreme ends by specific but often correlated risk perception characteristics (Table 55.2). Lay and expert persons were surveyed for their per­ ception of 81 hazards, and the aggregated results could be qualified by its position within one of four quadrants. In Slovic’s (1987) analysis, pesticides were positioned as a dread/unknown risk, although considered compara­ tively less risky than radioactive waste. Interestingly, the pesticide DDT, which had been banned for a decade at the probable time of data collection, appeared lower in the dread/unknown quadrant than pesticides in general. Perhaps constantly hearing about a particular chemical product influenced respondents to rank it as more known, albeit still unacceptably hazardous. The rodenticide D-Con (warfarin formulation), which remains on the over-the-counter market today, ranked in the dread/known

quadrant but was positioned very close to the axes, sug­ gesting that homeowners were familiar with the prod­ uct and its associated hazards, which were manageable because they had control of using it. The positioning of a common household pesticide like D-Con also suggests that benefits can influence positioning of risk perception. On the other hand, the analysis suggests that just the word pes­ ticide conjures an uncontrollable technology wherein the consumer does not feel they receive any benefit. Although the Slovic (1987) analysis is now over 20 years old, perusal of contemporary newspaper headlines suggests that a new survey is unlikely to change the posi­ tioning of pesticides along the perception factor continuum. For example, a Pew Charitable Trusts-sponsored survey (1999) indicated that 86% of respondents thought environ­ mental factors such as pollution were an important cause of the increased rate of disease. These beliefs differed somewhat between political parties, as seen in the survey of state legislators (Cohen, 1997), but at least 80% of all respondents agreed that pollution was a major etiologi­ cal factor. Over 80% of all respondents deemed pesticide

Factor 2 Unknown Risk

No Dread/Unknown Risk

Water fluoridation Saccharin

Nitrites Hexachlorophene Water chlorination Polyvinyl chloride

Dread/Unknown Risk DNA Technology

Nitrogen fertilizers Radioactive waste

Oral contraceptives Pesticides

Antibiotics

PCBs Mercury

Auto lead Lead paint

Caffeine Aspirin

DDT Fossil fuels Coal burning

Vaccines Disease caused by smoking

Auto exhaust D-Con (rodenticide)

Factor 1 Dread Risk Nerve gas accidents

Alcohol

Fires caused by smoking

Fireworks

NO Dread/Known Risk

Dynamite

Dread/Known Risk

Figure 55.1  Risk perception graph showing chemically related hazards qualified on a two-factor continuum of dread and unknown risk (Modified from Slovic, 1987).

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Table 55.2  Two-Factor Risk Perception Analysis, Showing Characteristics Associated with Each Direction of the Factors Factor 1 (dread risk)

Factor 2 (unknown risk)

High end

Low end

High end

Low end

Dread

Not dread

Unknown to those exposed

Known to those exposed

Uncontrollable

Controllable

Not observable

Observable

Global catastrophic

Not global catastrophic

Effect delayed

Effect immediate

Not equitable

Equitable

New risk

Old risk

Catastrophic

Individual

Risks unknown to science

Risks known to science

High risk to future generations

Low risk to future generations

Not easily reduced

Easily reduced

Risk increasing

Risk decreasing

Involuntary

Voluntary

Based on Slovic (1987).

residues in food as having a great impact on public health, but 91% thought we needed to learn a lot more about the health effects of pollutants. In another study, purchasers of nonorganic produce rated the median annual fatality rate due to pesticide resi­ dues in the range of annual mortality risk from motor vehicle accidents in the U.S. (Williams and Hammitt, 2001). Surprisingly, pesticide residues were rated as more hazardous than food-borne microbial pathogens, although the latter were certainly discussed in the general media as producing definitive adverse health effects. Trust in gov­ ernment agencies charged with overseeing food safety was inversely correlated with higher perceived risks from eat­ ing conventional produce. Reduction of pesticide residues associated with organic agriculture influenced affirmatively the benefits index, suggesting that consumers tend not to perceive personal benefits from farmer’s use of pesticides. A comparatively higher level of education was the only demographic factor that predicted relatively lower risk per­ ceptions about pesticide residues. The inverse correlation between educational level and risk perception of pesticides was confirmed in a differ­ ent study that investigated consumer attitudes about the acceptability of foods grown with fertilizers, pesticides, preservatives, and genetic engineering (Moe et al., 2001). Pesticides were less of a concern (21%) to respondents than genetically modified foods (49%). Age of respond­ents was another influential factor trending positively toward increased acceptability of pesticides, fertilizers, and genetic engineering. Ability to understand quantitative relationships and probability likely influences risk perception. The term innumeracy has been coined to describe “an inability to

deal comfortably with the fundamental notions of number and chance” (Paulos, 1988). The risk perception and com­ munication literature over the last decade has increasingly addressed numeracy, with one definition presented as “the ability to understand and manipulate proportions, risks, percentages, and probabilities” (Dieckmann et al., 2009). Perhaps pejoratively described as mathematical illiteracy, innumeracy describes a condition of misperception of the magnitude of numbers and ignorance of the mathematics of probability (Paulos, 1988). That such a state of compe­ tency exists is ironic in a society whose media outlets dis­ play daily public opinion polls with stated error rates. The need to now examine numeracy among audiences needing risk information stems from findings that how information is displayed can influence decision making. Attention has been particularly focused on professionals who must make medical decisions to reduce risk or who must communicate with patients to help them understand the magnitude of risk (e.g., Galesic et al., 2009; Gigerenzer and Edwards, 2003; Hoffrage et al., 2000). The implications of the research is that humans have intrinsically more diffi­ culty comprehending probabilities of events or effects than they do natural frequencies of them (Butterworth, 2001). Other research suggests that a narrative accompany­ ing risk information may make use of a percentage like­ lihood (i.e., a probability) format feasible (Dieckmann et al., 2009). Lay decision-maker subjects scoring higher in numeracy tended to rate a risk lower compared to sub­ jects scoring lower in numeracy. An example of a narrative for pesticide-associated risk might be qualitative informa­ tion about specific adverse effects suffered by subject test animals at any dose. A percentage or likelihood perspec­ tive would focus on the observation that only 15% of the

Chapter | 55  Communicating Safe Pesticide Use

animals tested reacted adversely. In experiments testing the interaction of the narrative, likelihood assessment, and numeracy, less numerate subjects relied more on a narra­ tive to make a decision about risk. Thus, people perceive risk through narratives, and understanding of probabilities of risk cannot be expected from consumers unless a certain level of numeracy exists. In sum, pesticides have been and continue to be per­ ceived as dreaded and unfamiliar risks. Such perceptions are unlikely to change despite the advent of compounds that are decidedly less hazardous than ever. The factors affecting risk perception of pesticides (as well as other technologies) based on over 30 years of research can be summarized in the following list. Consumers generally do not understand the technology (i.e., beyond simply killing pests). l Risks are involuntary because someone else is using them, but you may be exposed. l Similarly, you have a low degree of control and involve­ ment in use. l The threat to one’s self is considered great (or likely). l There is a lack of trust in decision makers. l Information about hazards is contradictory (likely exac­ erbated by the accessibility of the Internet as a reposi­ tory of any and all information). l Benefits are not clear (i.e., seen to accrue to users of pesticides but not to oneself). l

55.5  Perceptions of risk: who believes what and whom? An updated presentation of the two-factor dimensional analysis of risk perception showed how experts per­ ceived risk attributes of pesticides relative to nonexperts (Fischhoff, 2009). Experts tended to also place pesticides in the dread/unknown quadrant but much closer to the graph axes, suggesting less feeling of dread than nonexperts. Although early research as well as intuition would sug­ gest that experts and lay people have large differences in how risks are perceived, further research showed that such

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common wisdom is not predictive of actual perceptions. Indeed, the research shows that there are divergent differ­ ences among toxicologists themselves (Kraus et al., 1992; Mertz et al., 1998; Slovic et al., 1995). First, differences of opinion occur among toxicologists depending on whether they are affiliated with academia, industry, or regulatory agencies (Figure 55.2). Toxicologists as a whole seem to disagree with the lay public depending on the question posed, but sometimes agreement is remarkably similar (Figure 55.3). In sum, thinking that scientific illiteracy hampers risk communication ignores the diversity of opinions within the expert community about toxicological principles that one would assume had universal agreement within the dis­ cipline. If trust and credibility (or lack thereof) are major factors influencing effective risk communication, then the adage of “cleaning one’s own house first” might be applic­ able when crafting a risk message. Risk communication research shows a definitive relationship between acceptance of delivered messages about a subject and trust in the person who is presenting the information. Although the public seeks information, experts are not necessarily received in a friendly man­ ner. In referring to a seemingly mistrustful audience, one researcher has coined the phrase “decline of deference” (Laird, 1989), defining it as “a situation in which a hos­ tile and alienated public is mobilized primarily through ad hoc voluntary organizations, and is increasingly reluc­ tant to defer important decisions to institutional elites.” The phrase was applied to the problems communities have with siting unpopular facilities such as hazardous waste incinerators, nuclear power, or radiological waste reposi­ tory plants. However, any “expert” who has had to speak to an audience about a pesticide-contaminated well affect­ ing a community will attest that the decline of deference is far-reaching. Coincident with increased public distrust of formerly trusted institutions has been the rise in advocacy or special interest groups, including environmental organi­ zations. Indeed, with the growth of the Internet, environ­ mental advocacy groups have become major sources of information about chemical hazards along with the EPA Figure 55.2  Diversity of expert opinion as recorded by percentage response to statement. (A) There is not safe level of exposure to a cancercausing agent. (B) Our society has perceived only the tip of the iceberg with regard to the risks associ­ ated with chemicals. Graphs constructed from data presented in Kraus et al. (1992).

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Figure 55.3  Comparison of lay public and toxicologists response to the following questions: (A) If you are exposed to a toxic chemical substance, then you are likely to suffer adverse health effects. (B) For pesticides, it’s not how much of the chemical you are exposed to that should worry you, but whether or not you are exposed to it at all. (C) The way that an animal reacts to a chemical is a reliable predictor of how a human would react to the same chemical. (D) If a scientific study produces evidence that a chemical causes cancer in animals, then we can be reasonably sure that the chemical will cause cancer in humans.

and the CDC (specifically the Agency for Toxic Substance and Disease Registry). Evidence for the decline of deference is as germane in 2010 as it was 20 years ago, including a preference for independent political affiliation, a decline in prestige for formerly venerated occupations such as physician and teacher, loss of confidence in various institutions ranging from medical establishments to Congress, and a drop in voter turnout. With regards to chemical technology, decline of deference is surely pushed by arguments that risk assess­ ment is deeply flawed and does not adequately protect the public or environment (Buhl-Mortensen, 1996; Schorling, 2003). After all, every regulation and decision made by the EPA is fought over in public through the courts and under the scrutiny of the media, and scientific experts take every side imaginable (Neil et al., 1994). Loss of trust in risk assessment is likely further engendered by reports of a plethora of chemicals residing in our bodies as well as everywhere in the environment and wildlife. The gradual building of public distrust in governmental institutions is also coincident with messages that a flawed governmental policy, such as risk assessment, has been negatively per­ ceived. Is it not surprising, therefore, that the breach has been increasingly filled with calls for a new paradigm

called the precautionary principle. Thus, any attempt to communicate about the safe use of a technology will have to convince a distrustful public that the “broken” policy of risk assessment, whether true or not, has been replaced with something new and improved. Furthermore, the loss of deference means that risk communicators cannot simply swamp the public with information and numbers. Demands are for open dialog and transparency. Those who come off as trustworthy and empathetic will find it easier to attain credibility in a mistrustful environment. The apparent loss of deference does not mean that trust and credibility cannot be re-established or at least sought out. Surveys have shown that three sets of determinants account for a significant amount of variation in perceptions of trust and credibility: knowledge and expertise, honesty and openness, and concern and care. Thus, risk commu­ nicators exhibiting these qualities or characteristics are more likely to be listened to with a somewhat open mind. Hypothesis testing showed that the importance of the afore­ mentioned three sets of factors in increasing perceptions of trust and credibility varied by the group doing the communi­ cating (Peters et al., 1997). For example, industry represen­ tatives showing concern and care garner greater increases in perceptions of trust and credibility. A perception

Chapter | 55  Communicating Safe Pesticide Use

Table 55.3  Response of Washington State Households to the Question “In Your Opinion, Is [Topic Inserted] a Problem in Your Community?” Topic

% Yes

% Somewhat % No

Outdoor air quality

13.3

9.0

76.6

Drinking water quality

11.1

4.4

82.3

Workplace hazards

  9.5

4.1

85.5

Solid waste management

  7.3

2.7

88.2

Pesticide use and control

  7.1

2.8

84.3

Wastewater management

  7.0

2.3

86.8

Hazardous waste sites

  6.3

1.6

86.8

Air quality inside home

  2.6

2.6

93.6

Data from Laflamme and VanDerslice (2004).

of increased commitment by government officials and increased knowledge and expertise by citizen groups gar­ nered increases in trust and credibility. Finally, we should also be wary of forcing too much information upon an audience that is not really looking for it. One study (Cohen, 1997) definitively shows that legislators and staff are seeking more information about all aspects of risk. However, surveys of consumers sug­ gest that pesticides are not a spontaneous issue (U.K. Food Standards Agency, 2003). In other words, until confronted with having to think about a specific pesticide issue, most consumers just may not be interested. This conclusion was also suggested in a survey of Washington State communi­ ties using the Behavioral Risk Factor Surveillance System (BRFSS) (Laflamme and VanDerslice, 2004). The survey instrument questioned households about their concerns over various environmental problems, including pesticide use (Table 55.3). Only 10% of household respondents viewed pesticides as an environmental problem in their community, whereas 22% rated outdoor air quality of more concern. If the Washington State BRFS survey results are representative of other state surveys, then perhaps crop and public health protection managers can take the time to seek information and raise the probability of successfully com­ municating their viewpoints while pesticides are not “on the front burner of consumer concerns.”

55.6  Needed information: some ideas for presenting risk concepts Risk communication research overwhelmingly supports determining what knowledge or beliefs an audience already has prior to crafting a message (Morgan et al.,

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1992, 2002). Lay persons will come to a meeting, for example, with some knowledge about a subject, but it is not the same technical knowledge that the risk communi­ cator is aware of. Rather, the knowledge is more akin to a belief system that has developed in an individual from either direct access of information on their own, from the media or advocacy groups, or through their networks of acquaintances. Such beliefs are a form of mental model that will be relied upon to make a decision about how risky pesticides, or any other technology or activity, are. Simply inundating people with information that an expert deems useful is unlikely to achieve a goal of understanding and/or making favorable decisions about a technology. Thus, lis­ tening to the audience and gathering concerns is arguably the most important step in initiating a risk communication effort. Understanding the mental models held by an audi­ ence will thus require efforts at risk communication early on, rather than as an ad hoc process once a pest manager has already made a decision to spray a large swath of pub­ lic land for noxious weeds. Even before gathering information about the public’s various mental models, an expert’s mental model should be delineated (Morgan et al., 2002). If the experts are perceived to have an uncertainty themselves about their subject and a dubious understanding of their own ideas about risk of a chemical technology, than the public can­ not be expected to be open to a change in their percep­ tions. Furthermore, elaboration of an influence diagram as part of this process will help to organize the network of parameters or factors driving a risk characterization pro­ cess and should therefore engender a better understanding of how one factor influences another factor (Morgan et al., 1992, 2002). For example, an influence diagram for the likely exposure of bystanders to a roadside herbicide spray might start with a circle representing the sprayer itself. The sprayer circle would then have an arrow pointing to several other circles that might represent airborne aero­ sols, plant surfaces, water, etc. Such receptors of pesticide spray residues are likely to be present in any application area. Thus, the nature of the sprayer directly influences the generation and interception of residues. From the recep­ tors themselves, arrows could extend specifically to peo­ ple coincidentally walking in the sprayed area, people in nearby homes, or pets playing outside. In these compart­ ments that represent the organismal receptors, residues would be absorbed, metabolized, etc. Further relationships between dose at the body surface and dose absorbed would be delineated to specific hazards. The bottom line is that one would basically end up with a network of exposure and effects processes represented in a hierarchical fashion. The value in any one process factor is directly influenced by the factor pointing to it. Pest control managers are likely aware that the laws that regulate pesticides and the mandated processes for safety testing, registration, and labeling can be somewhat

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daunting and confusing regarding their interrelation­ ships. Simply dismissing everything as FIFRA (Federal Insecticide Fungicide Rodenticide Act) misses the fact that parallel laws exist (the partner law being the FFDCA, Federal Food Drug and Cosmetic Act) and the relation­ ship between risk assessment, registration, and labeling. However, this information is perhaps most crucial to be intimately familiar with so that the audience for risk mes­ sages at least has the opportunity to modify a common mental model that the government has little control over pesticide technology. The influence diagram could be one way to visualize how the laws are interrelated. I use such a model in a slide called “Pesticide Law 101” dur­ ing extension training programs (Figure 55.4). Specific elements under “Risk Assessment” could be further expli­ cated by relying on the standard risk characterization and decision-making paradigm adopted by the EPA Office of Pesticide Programs. One important aspect of this para­ digm is that it has been vetted over the last 25 years, espe­ cially by the National Academy of Sciences (NRC, 1983, 1993, 1994). In the mental model paradigm for risk communication, the process of obtaining the necessary background from public stakeholders could be arduous and time-consuming. For example, the strategy for determining extant beliefs among the audience for risk communication would involve a four-step process beyond constructing the expert model (Morgan et al., 2002): Conduct mental models interviews to elicit beliefs about a hazard expressed in a lay person’s terms. l Conduct structured initial interviews to confirm lay person’s beliefs and the expert model. l Draft a risk communication to determine which beliefs about a risk need correcting or have gaps requiring fur­ ther information. l Evaluate the communication to ensure that “audience members” understand it as intended. l

The above paradigm for successful communication is idealistic in that the process assumes a sufficient amount of time and resources to gather background information about public beliefs and test out communication messages. However, the time and resources available will depend on the context of need for risk communication. For example, if a natural area burned by a fire will need revegetation management, and a herbicide is planned to be one tool in the process, sufficient time and funding might be available if the activity was not going to occur for many months after the fire. Similarly, an ongoing program of roadside spray­ ing has the time to work with citizens that might perceive harm from the program. On the other hand, a drift inci­ dent wherein harm is perceived as imminent or has already happened may require throwing together a message fairly quickly. Under these various scenarios, I would argue that enough information has been garnered about public

FIFRA (1947)

FFDCA (1938)

FEPCA (1972)

Risk Assessment

Miller (1954) Delaney (1958)

Tolerance (“MRL”)

Labeling

Registration

FQPA (1996)

Figure 55.4  Pesticide Law 101, a modified type of influence diagram to show the interrelationships in one simplified picture among the parallel laws that control pesticide safety assessments on the way to product regis­ tration and labeling. In this case, the arrows represent the controlling laws and the processes involved. Miller, Delaney, FEPCA, and the FQPA are all amendments that authorized new mandates to the parallel main stat­ utes. (FIFRA, Federal Insecticide, Fungicide, Rodenticide Act; FFDCA, Federal Food Drug and Cosmetic Act; FEPCA, Federal Environmental Pesticide Control Act; FQPA, Food Quality Protection Act; MRL, Maximum Residue Limit).

perceptions of pesticides since the heyday of DDT to make some generalities about what mental models people hold and thus what information is needed to reorient mental models or fill in knowledge gaps. For example, I first make the simplistic argument that those expressing concerns about the use of pesticides in a community often know something about pesticide hazards from information garnered on the Internet, and perhaps more importantly from the many websites hosted by envi­ ronmental advocacy groups. Thus, the first misconception to overcome is the need to differentiate hazard and risk. Without this fundamental differentiation at the start, I do not think it possible to successfully convey why a pesti­ cide application’s benefits can be far greater than its risks, because risks and hazards, obtained by two different pro­ cesses will be conflated. The National Research Council (NRC) has defined hazard as “An act or phenomenon posing potential harm to some person(s) or thing(s); the magnitude of the hazard is the amount of harm that might result, including the seriousness and the number of peo­ ple exposed” (NRC, 1989). A report by the International Program on Chemical Safety/Organization for Economic Cooperation and Development (IPCS/OECD) Working Party, has offered the following definition for hazard: “inherent property of an agent or situation capable of hav­ ing adverse effects on something” (Duffus, 2001). One could find many similar definitions for hazard by a casual search on the Internet, but all simplify to the concept of any technology or event possessing potential harm. The NRC (1989) defined risk as “adds to the hazard and its magnitude the probability that the potential harm or undesirable consequences will be realized.” The IPCS/ OECD (Duffus, 2001) defined risk as “the probability of adverse effects caused under specified circumstances by an

Chapter | 55  Communicating Safe Pesticide Use

agent in an organism, a population, or an ecological sys­ tem.” Thus, all definitions of risk simplify to the concept of likelihood or probability that something bad will hap­ pen. Importantly, however, the probability of the hazard manifesting itself (i.e., becoming more than just potential harm) is situational or conditional. My second observation about common misperceptions stems partly from the confusing of hazard and risk and partly from the plethora of past and present news reports about the ubiquity of pesticide residues in the environment. Thus, citizens are unlikely to believe that mere detection of a pesticide residue does not equal a hazard and expo­ sure to the residue is unlikely to be without consequence. These misconceptions can be hard to overcome because they have to be considered as probabilities (or likelihood) of adverse events given a certain exposure. Yet risk com­ munication research warns against expressing messages as probabilities rather than as frequencies. Perhaps a way around this dilemma is to confront head on just what infor­ mation is available. Thus, an examination of registration eligibility decision documents (REDs) on EPA’s web­ site will offer for various pesticides the doses used to set toxicity benchmarks. If one honestly conveys such narra­ tive information to the public, I believe you set the stage of trust and credibility (i.e., you are not holding anything back). As trust and credibility builds you can then ease the audience toward seeing that such knowledge was garnered only after an impossible exposure situation (barring the issue is a direct spill of a product on the skin). Inevitably, the discussion must focus on the fact that such high haz­ ards are informative but the real useful information is what does not occur at the exposure dosage equivalent to the NOAEL (no observable adverse effect level). At this point, one could more facilely discuss what likely exposures are in an environmental context. Regarding the latter discussion of exposures with the public, temptation will lead to the inevitable exposure, hazard, or risk comparison. A word of caution is wise here because communication research shows that such compari­ sons must be for phenomena of comparable qualities. For example, if chemical exposure is the issue, then any com­ parisons must be to chemicals that the public is likely to have personal experience with. Such examples can be found by examining the toxicological information available for “common household chemicals,” including commonly used over-the-counter medications such as aspirin or acetamino­ phen or perhaps caffeine, commonly consumed in soft drinks in addition to coffee. Nicotine is a particularly toxic household chemical if one considers that approximately 22% of the U.S. population are smokers (Healton et al., 2006). However, nicotine is also associated with a lot of negative messages about tobacco use and associations with lung cancer. By using something easily recognized (yet still considered chemical in nature), one can also use a pic­ ture to illustrate just how much test rodents are exposed

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Figure 55.5  Expression of 2,4-D daily rodent dosage as units of aspirin. The three dose levels were given ad lib via diet to rodents in a chronic toxicity study. Each dose was scaled to the body weight of a human female and then pelleted into 325-mg/kg aspirin tablets. (Based on Charles et al., 1996; NOAEL was determined to be 5 mg/kg/day with no evidence of oncogenicity.)

to on a daily basis, for anywhere from 90 to 700 days, to understand the array of possible adverse effects. By scal­ ing the dosages of 2,4-D administered daily to rodents in one study to the body weight of a female adult and then pelleting the dose as aspirin tablets, I have used a picture to illustrate just how large a daily dose of anything is at the milligrams per kilogram level of exposure (Figure 55.5). If possible, use familiar chemicals that people know are beneficial at one dose but potentially hazardous at high doses because this same counterintuitive concept could be applied to pesticides. A third type of misconception that I have detected is a feeling that the EPA registers pesticides without thor­ ough knowledge about their risk to the environment and/ or people. This misconception probably relates to the lack of trust/credibility issue raised in risk communication research. Certainly just reading Internet advocacy sites might suggest that little is known about specific pesticides. In surveying the landscape of chemical regulatory law, it is true that the vast majority of chemicals released into com­ mercial use have not had the scrutiny that pesticides have had. But communication must fill the gap in the mental model that all chemicals, including pesticides, suffer from a deficiency of scrutiny. At this point, one could enumerate the countless tests that regulatory law requires specifically for pesticide technology while acknowledging that the law for control of other chemicals is not nearly as intensive in requirements. Thus, one strategy is to separate as clearly as possible the unique regulatory status of pesticides in com­ parison to other chemicals. As an adjunct to this, I often

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show pictures from the EPA website that detail all the tests required prior to registration. Furthermore, I describe the mandatory process of conducting all studies under Good Laboratory Practice requirements, emphasizing that all data is subject to independent auditing. With regard to data collection, I point out that one has to recognize the objectives of studies that EPA relies on to make their decisions about safety of pesticides. I am doubtful that even pest management specialists know the difference between a study conducted to learn about mech­ anisms of a toxic effect versus studies conducted to under­ stand what the possible array of effects are and establish a threshold of effect. From the perspective of understanding risk, the mental model regarding the validity of all toxicity studies should be influenced toward an understanding that only certain types of studies are predictive of what would happen in the environment. Again, communicating about this issue must be presented without seeming to hide data. Thus, a side-by-side comparison to show people what is entailed in the two different kinds of studies may be well received. One aspect of information needs for potentially chang­ ing mental models and filling knowledge gaps pertains to a seemingly growing confusion with risk assessment and risk management. Each endeavor as applied to pesticide tech­ nology does not stand alone. A bridge exists between them in that one informs the other. However, how these endeav­ ors are conducted is different in approach and consider­ ations of data needs. For example, the risk characterization part of risk assessment entails standard hypothesis testing within the context of scientific practices. Discussions of hazard identification, dose–response testing, and exposure assessment provide an opportunity to review the concept of science as a process to find out information as opposed to a collection of facts. On the other hand, the public needs to know that the final risk characterization is part science pro­ cess and part policy process, which falls squarely into the camp of risk management. Integration of expected expos­ ure and some toxicological endpoint is pure mathemat­ ics, albeit construction of a simple ratio. But application of an uncertainty factor (preferably in nonjargon terms, a safety factor) and a decision on what the magnitude of the ratio means are purely policy decisions falling under risk management. I suspect that often confusion and per­ haps conflict over different interpretations of a risk assess­ ment occur because the participants overlook what part of the process they are dealing with (Figure 55.6). How can the public not be equally conflicted if experts forget the fine line between risk characterization and management. Furthermore, we should remember that defining the array of tests that are expected by a regulatory agency from a manufacturer is in itself a form of risk management, i.e., the tests are mandated by science policy. Finally, regarding the potential confusion between characterizing risks and managing them, raises the issue

Hayes’ Handbook of Pesticide Toxicology

Experimentation (measurement, data analysis) Scientific Risk Assessment Socio-political Economic Risk Management Figure 55.6  The relationship between risk management and risk assessment and the associated influence of process (i.e., scientific endeavor vs. sociopolitical–economic considerations) on each. Risk man­ agement definitively influences risk assessment through science policy (e.g., choice of tests) and risk assessment influences risk management by presentation of data from risk characterizations. However, risk assessment is not a pure scientific process because the magnitude of acceptable risk (which is simply a ratio of estimated exposure to toxicological bench­ mark) is determined by policy.

of advocacy for the precautionary principle in lieu of risk assessment. The validity of the substitution is suspect because the precautionary principle is a conceptualization of a philosophical outlook, but risk assessment is a definitive set of procedures used to characterize risk. Communicators crafting risk messages about pesticides should be aware that a public mental model may be forming around a pref­ erence for the precautionary principle. Thus, when asked in a public meeting, the informed pesticide risk communicator should be familiar enough with the concept to show a dis­ trustful audience that not only has the idea been considered, but pertinently upon deeper scrutiny the current system of pesticide registration that has evolved over a long time is in fact the precautionary principle in practice! The latter statement may seem blatant considering the historical conflicts over pesticide use initiated even before Silent Spring (Carson, 1962), but consider my argument of why such a message regarding pesticides is valid. First, the precautionary principle has not been rigorously defined except as a philosophical goal: “when an activity raises threats of harm to human health or the environment, precau­ tionary measures should be taken even if some cause and effect relationships are not fully established scientifically” (Kriebel et al., 2001). Attempts have been made to put the concept into a more scientific mode with respect to pro­ tecting the marine environment (Gray and Brewers, 1996). “The Precautionary Principle involves precautionary action to safeguard the marine environment by preventing and reducing emissions of hazardous substances at source and minimizing physical disturbance caused by human activi­ ties using appropriate technologies and measures. The Precautionary Principle shall apply to human activities for which there exists a scientific basis for believing that damage to habitats or harmful effects on marine species

Chapter | 55  Communicating Safe Pesticide Use

are likely to result. Measures shall be based on pessimis­ tic assumptions regarding uncertainties in the measurement and prediction of effects on the environment.” The former definition lays out no specific strategy other than to “use precaution,” but some attempt to show how it might work in stepwise fashion has been made available in an essay posted to the Internet (O’Brien, undated). Thus, part of the problem with a mental model favoring the precautionary principle is that the idea itself lacks clarity and consistency in use (Goldstein, 2005). Essentially, the scrutiny focused on defining and refining risk assessment has not been given to the precautionary principle. Hardly anyone could be opposed to taking precautions, but the principle is not likely to help make a decision about whether to use a technology or not. The latter point follows from the stated objective of precautionary principle advo­ cates: alternatives should be sought for substances posing “threats of harm to human health or the environment.” Yet, how does one know about the validity of the alternatives without a formal scientific assessment, or in other words a risk assessment? Here I argue that in fact, the long evo­ lution of the process for pesticide risk assessment, both codified into regulatory law as well as in the research beyond what is entailed for registration, is the epitome of the precautionary principle, especially after passage of the FQPA. For consumer protection at least, pesticide regula­ tions moved from a risk–benefit basis to one of consid­ ering risks only. Such a move is definitively in keeping with the objective of considering a chemical guilty until proven harmless (which is actually impossible to defi­ nitely prove). In the inevitable uncertainty associated with any risk characterization, the margin of exposure (MOE) approach applies an uncertainty factor of 100 to 1000 to ensure exposure would be orders of magnitude lower than NOAELs. Thus, uncertainty is not ignored but accounted for directly. The bottom line for risk communication is that pest managers need not ignore public calls for application of the precautionary principle but should communicate that in fact pesticide technology has long ago embraced it formally.

55.7  Benefits information: the most needed piece of the communication puzzle The foregoing section described some specific issues and possible ways to illustrate in a more accessible manner some aspects of pesticide risk assessment. The information was oriented to helping the public overcome some misun­ derstandings or gaps in information. What was presented would not necessarily help anyone make a decision about what is an acceptable risk regarding pesticide technology. But at least the information would help bring the public and experts to a more common ground of understanding.

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Decades of research has provided numerous guidelines for communicating about risk, and much of it could be use­ ful to those who must craft messages about pesticide use (e.g., general “how to” type information can be found in papers by Covello, 1998; Fischhoff, 1995; Morgan et al., 1992; Peterson and Higley, 1993; Rowan, 1994). One com­ mon element necessary for risk communication is benefits information. The public often feels that they do not receive any benefits from a technology that has been imposed on them and out of their control. Too often pesticide risk com­ munication boils down to talking about acceptable levels of risk and ignores what should be the good news: clear benefits. Not surprisingly, overlooking messages of ben­ efits may be coincidental with nearly all funding going to understanding hazards and risks and little to examining more quantitatively the benefits that accrue beyond the pesticide user. Discussion of benefits can be brought to situations that most of the public concerned with natural resources is likely to understand: invasive species. Out of control weeds vex every homeowner, but weeds can quickly over­ whelm natural areas that animals forage in. When talking about the benefits of herbicide use, one can emphasize that other options are usually weighed but found ineffective given the rapid spread of invasive vegetation. Recently, a couple of publications have attempted to make a stronger case for the benefits of pesticide technology (Cooper and Dobson, 2007; Gianessi and Reigner, 2007). Researchers have even shown that in the case of a disease such as West Nile Virus, the risk of contracting and suffering ill effects of the mosquito-borne disease outweighs any risk from the insecticide most likely to be used to control the vec­ tor (Peterson et al., 2006). The latter story is also illustra­ tive of the IPM approach to public health with widespread resources going into bird reservoir monitoring and mos­ quito population monitoring prior to taking any action. Such activities would be looked on favorably by precau­ tionary principle advocates, as illustrated in their assess­ ment of how pesticide use in schools could be curtailed (Kriebel et al., 2001). One final message that is best interpreted under the rubric of benefits information is the notion that pesticide technology and the relevant regulatory law and policies are dynamic. The whole technology has learned this from its overreliance on chemicals that were not compatible with environmental health and not used within the context of integrated pest management strategy. As more new test­ ing requirements and methodologies have been incorpo­ rated into the data analysis phase of risk assessment, new compounds that are arguably much safer are being rapidly commercialized. A few “old technology” compounds hang on but the overwhelming message to the public should be one of change. It cannot hurt to let the public know that pesticide technology has so completely changed over the last two decades that new products are winning Presidential

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Green Chemistry Awards whose purpose is to recognize sustainable and safer technologies. Indeed, how ironic that a company like Dow Agrosciences was one of the first to study, develop, and market spinosad, a compound that has been formulated into products acceptable for use in certi­ fied organic production as well as in other agricultural and urban sectors in general.

Conclusion Risk communication research points to common problems when conveying information about technological hazards. General factors involving risk perceptions of the lay per­ son include subject complexity, innumeracy, loss of con­ trol, decline of deference in experts and institutions, and lack of seeing benefits. Tendencies to blame the media should be dropped in favor of the understanding that the public is seeking information rather than a source for how to feel about a hazard. The nonexpert has a mental model that they bring to the table, so the risk communicator must strive to understand what it is and thereby work to change perceptions. At the same time, experts have mental models also, and research shows they do not necessarily agree with one another. Such uncertainty reinforces public distrust and lack of confidence. Regarding communicating about pesticide technology specifically, I hold that after decades of use, problems, and communication efforts, enough is known to understand common mental models. Some of these include confusion between hazard and risk, a feeling that the technology has not been studied sufficiently, and the impression that ben­ efits are absent. Several ideas were presented to overcome these perceptions. Given that the precautionary principle is fast becoming the model that should substitute for risk assessment, one overall strategy to communicate safe pes­ ticide use is to recognize that the technology as well as its regulation is very dynamic and proactive, taking action to reduce uncertainty. Even though the precautionary princi­ ple remains a philosophical concept rather than a roadmap of how to assess technology, pesticide risk communicators ought to use the concept to show in fact that deployment of the technology may indeed be the epitome of its idealism.

References Ames, B. N. (1992). Pollution, pesticides, and cancer. J. Assoc. Offic. Anal Chem. Intl. 75, 1–5. Buhl-Mortensen, L. (1996). Type-II statistical errors in environmental science and the precautionary principle. Marine Pollution Bulletin 32(7), 528–531. Butterworth, B. (2001). Statistics: What seems natural? Science 292 (5518), 853–855. Carson, R. (1962). “Silent Spring,” Houghton Mifflin Co., New York, p. 304������������������.

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Charles, J. M., Bond, D. M., Jeffries, T. K., Yano, B. L., Stoptt, W. T., Johnson, K. A., Cunny, H. C., Wilson, R. D., and Bus, J. S. (1996). Chronic dietary toxicity/oncogenicity studies on 2,4-dichloropohen­ oxyacetic acid in rodents. Fundamental and Applied Toxicology 33, 166–172. Cohen, N. (1997). The politics of environmental risk: Perceptions of risk assessment in the state legislatures. Policy Studies Journal 25(3), 470–484. Cooper, J., and Dobson, H. (2007). The benefits of pesticides to mankind and the environment. Crop Protection 26(9), 1337–1348. Covello, V. T. (1998). Risk perception and communication. Donna Buckley, Editor North American Conference on Pesticide Spray Drift Management March 29–April 1, 1998, Holiday Inn By the Bay, Portland, Maine; University of Maine Cooperative Extension: 161–186. Dezenhall, E. (1999). “Nail ‘Em!: Confronting High-Profile Attacks on Celebrities & Businesses,” Prometheus Books,���������������������������� p. 293���������������������. Dieckmann, N. F., Slovic, P., and Peters, E. M. (2009). The use of nar­ rative evidence and explicit likelihood by decisionmakers varying in numeracy. Risk Analysis 29(10), 1473–1488. Duffus, J. H. (2001). Risk assessment terminology. Chemistry International 23: http://www.iupac.org/publications/ci/2001/march/risk_assessment. html#29. Dunwoody, S. (February 1991). Do people really learn about risk from the mass media? In “Environmental Risk Reporting: The Science and Coverage” (S. M. Friedman and C. L. Rogers, eds.), pp. 25–27. Lehigh University. Fischhoff, B. (1995). Risk perception and communication unplugged: Twenty years of process. Risk Analysis 15(2), 137–145. Fischhoff, B. (2009). Risk perception and communication. In “Oxford textbook of public health” (R. Detels, R. Beaglehole, M. A. Lansang, and M. Gulliford, eds.), 5th ed., pp. 940–952. Oxford University Press, Oxford, England. Fischhoff, B., Slovic, P., Lichtenstein, S., Read, S., and Combs, B. (1978). How safe is safe enought? A psychometric study of attitudes towards technological risks and benefits. Policy Sciences 9, 127–152. Friedman, S. M., Villamil, K., Suriano, R. A., and Egolf, B. P. (1996). Alar and apples: newspapers, risk and media. Public Understanding of Science 5(1), 1–20. Galesic, M., Gigerenzer, G., and Straubinger, N. (2009). Natural frequen­ cies help older adults and people with low numberacy to evaluate medical screening tests. Medical Decision Making 29, 368–371. Gianessi, L. P., and Reigner, N. P. (2007). The value of herbicides in U.S. crop production. Weed Technology 21, 559–566. Gigerenzer, G., and Edwards, A. (2003). Simple tools for understand­ ing risks: from innumeracy to insight. British Medical Journal 327, 741–744. Goldstein, D. A. (2005). The precautionary principle: Is it a threat to toxi­ cological science? International Journal of Toxicology 25, 3–7. Gray, J. S., and Brewers, J. M. (1996). Towards a scientific definition of the precautionary priinciple. Marine Pollution Bulletin 32(11), 768–771. Griffin, R. J., Dunwoody, S., and Zabala, F. (1998). Public reliance on risk communicaiton channels in the wake of a cryptosporidium out­ break. Risk Analysis 18(4), 367–375. Healton, C. G., Vallone, D., McCausland, K. L., Xiao, H., and Green, M. P. (2006). Smoking, obesity, and their co-occurrence in the United States: cross sectional analysis. British Medical Journal 333d, 25–26. Hoffrage, U., Lindsey, S., Hertwig, R., and Gigerenzer, G. (2000). Communicating statistical information. Science 290(5500), 2261–2262. Kraus, N., Malmfors, T., and Slovic, P. (1992). Intuitive toxicology: Expert and lay judgements of chemical risks. Risk Analysis 12(2), 215–233.

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Communicating Safe Pesticide Use

Kriebel, D., Tickner, J., Epstein, P., Lemons, J., Levins, R., Loechler, E. et al. (2001). The precautionary principle in environmental science. Environmental Health Perspectives 109(9), 871–876. Laflamme, D. M., and VanDerslice, J. A. (2004). Using the behavioral risk factor surveillance system (BRFSS) for exposure tracking: Experiences from Washington State. Environ. Health Perspect. 112, 1428–1433. Laird, F. N. (1989). The decline of deference: The political context of risk communication. Risk Analysis 9, 543–550. Mertz, C. K., Slovic, P., and Purchase, I. F. H. (1998). Judgments of chemical risks: Comparisons among senior managers, toxicologists, and the public. Risk Analysis 18(4), 391–404. Moe, G., Neuhouser, M., Kristal, A., Rock, C., and Neumark-Sztainer, D. (2001). Consumer acceptability of fertilizers, pesticides, preservatives, artificial sweeteners, fat substitutes, and genetically modified foods. Journal of the American Dietetic Association 101(9 Supplement 1), A–40. Morgan, M. G., Fischhoff, B., Bostrom, A., Lave, L., and Atman, C. J. (1992). Communicating risk to the public. Environ. Sci. Technol 26, 2048–2056. Morgan, M. G., Fischhoff, B., Bostrom, A., and Atman, C. J. (2002). “Risk Communication: A Mental Models Approach,” Cambridge University Press, Cambridge, UK, p. 366. National Research Council (NRC) (1983). “Risk Assessment in the Federal Government: Managing the Process.” National Academy Press, Washington, DC. National Research Council (NRC) (1987). “Regulating pesticides in food: The Delaney paradox.” National Academy of Sciences, Washington, DC, p. 288. National Research Council (NRC) (1989). “Improving Risk Communication,” National Academy Press, Washington, DC. National Research Council (NRC) (1993). “A paradigm for ecological risk assessment,” pp. 241–201. In Issues in Risk Assessment. National Academy Press, Washington, DC. National Research Council (NRC) (1994). “Science and Judgment in Risk Assessment.” National Academy Press, Washington, DC. Neil, N., Malmfors, T., and Slovic, P. (1994). Intuitive toxicology: Expert and lay judgments of chemical risks. Toxicologic Pathology 22(2), 198–201.

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O’Brien, M. H. Beyond Democratization of risk assessment: An alternative to risk assessment. Science and Environmental Health Network, URL: http://www.sehn.org/conbiorisk.html (accessed Oct 4, 2009). Paulos, J. A. (1988). “Innumeracy: Mathematical illiteracy and its consequences.” Hill and Wang, New York, p. 186. Peters, R. G., Covello, V. T., and McCallum, D. B. (1997). The determinants of trust and credibility in environmental risk communication: An empirical study. Risk Analysis 17(1), 43–54. Peterson, R. K. D., and Higley, L. G. (1993). Communicating Pesticide Risks. American Entomologist 39(4), 206–211. Peterson, R. K. D., Macedo, P. A., and Davis, R. S. (2006). A human-health risk assessment for West Nile virus and insecticides used in mosquito management. Environmental Health Perspectives 114, 366–372. Pew Charitable Trusts (1999). “Public Opinion research on Public Health, Environmental Health and the Counry’s Public Health Capacity to Adequately Address Environmental Health Problems.” Johns Hopkins School of Hygiene and Public Health, Baltimore. Rowan, K. E. (1994). Why rules for risk communication are not enough: A problem-solving approach to risk communication. Risk Analysis 14(3), 365–374. Schorling, I. (2003). The Greens perspective on EU chemicals regulation and the white paper. Risk Analysis 23(2), 405–409. Slovic, P. (1987). Perception of risk. Science 236, 280–285. Slovic, P. (1991). Perceptions of risk: paradox and challenge? In “Environmental Risk Reporting: The Science and the Coverage Proceedings of a Workshop” (S. M. Friedman and C. L. Rogers, eds.), pp. 7–12. February 1991, Bethlehem, PA Lehigh University. Slovic, P., Malmfors, T., Krewski, D., Mertz, C. K., Neil, N., and Bartlett, S. (1995). Intuitive toxicology. II. Expert and lay judgments of chemical risks in Canada. Risk Analysis 15(6), 661–675. U.K. Food Standards Agency (2003). Consumer attitudes to pesticide use and pesticide residues. Consumer Committee Report ConsCommD028/03; URL: www.food.gov.uk/multimedia/pdfs/consattpesticideres.pdf (accessed October 2, 2009). Williams, P. R. D., and Hammitt, J. K. (2001). Perceived risks of conventional and organic produce: Pesticides, pathogens, and natural toxins. Risk Analysis 21(2), 319–330.

Section VIII

Regional and Global Environmental Exposure Assessments

(c) 2011 Elsevier Inc. All Rights Reserved.

Chapter 56

Ecotoxicological Risk Assessment of Pesticides in the Environment Keith R. Solomon University of Guelph, Guelph, Ontario, Canada

56.1  Introduction In general, risk assessment is a process of assigning magnitudes, probabilities, and relevance to the adverse effects that may result from a particular activity or set of activities. Ecotoxicological risk assessment (ERA), as we use it today, had its scientific origins in the area of human health risk assessment, but it has changed and adapted to address new questions, such as ecological consequences, and new approaches, such as probabilistic risk assessment (PRA). Human health risk assessment is focused on the protection of individual humans with a high level of certainty (see Chapter 84); however, ERA is mostly conducted with the objective or goal of the protection of populations and communities and their function in the environment. The exception to the population and/or community focus in ERA is when risks accrue to highly valued species or endangered organisms. In these cases, individual organisms will be protected in a similar way to humans. Therefore, in general, ecological risks are perceived by the public as being different from risks to individual humans. This is particularly true where there is a lack of identification with the organisms to which the risk accrues (Suter et al., 2007) as is illustrated in the difference between the acceptability of a risk that accrues to freshwater invertebrates (which will be wet and sometimes also slimy) and one that accrues to societally important organisms, such as pandas or threatened and endangered species what are protected by laws such as in the Endangered Species Act in the United States (FESTF, 2002). This is likely the basis for the inclusion of some animals (charismatic megafauna, usually fuzzy, anthropoid, and with eyes facing to the front!) in wildlife protection programs, while those with scales, mucous secretions, or too small to be seen with the naked eye are often relegated to lesser positions of importance. Thus, activities such as mining, urbanization, farming, and forestry are regarded as acceptable risks to the environment as Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

long as the organisms at risk are anonymous or nonanthropomorphic. Times change and so do social values. Frogs have only recently moved higher on the list of socially valued animals and are now afforded greater interest and protection than in the past (Mendelson et al., 2006). Because the focus of ecotoxicology is on the environment and the very diverse species and communities in the environment, it is strongly tied to ecology. While toxicology provides a description of the results of exposure to a stressor, ecology provides the basis for assessing the relevance of those responses in the larger context, that of the population, community, or ecosystem. Because of this strong link to ecology, it has been suggested that ERA is more properly defined as stress ecology (van den Brink, 2008; van Straalen, 2003), but it really is a combination of ecology and toxicology with ecology science providing the means to characterize the consequences of the stress and the toxicology providing the understanding of how the stress is initiated and how it propagates from the site of toxicological action to the population (Figure 56.1).

56.1.1  Risk Assessment of Pesticides ERA of pesticides is a special case of risk assessment applied to the use of upwards of 700 pesticide active ingredients, with a wide range of physical, chemical, and biological properties. Pesticides are chemical substances or physical processes deliberately used to control organisms (defined as pests) for the protection of crops, human health, or structures. Although physical agents, for example heat, microwave, or -radiation are sometimes used as “pesticides,” the focus of this chapter is on chemical pesticides. Use of pesticides to control pests always occurs after some form of risk assessment has taken place in relation to the particular pest problem. For example, the cost of the pesticide may be a considered in relation to the benefit resulting from the control of the pest. Thus, the cost of a 1191

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Ecosystem Ecology explains consequences

Population Organism Physiology Biochemistry

Toxicology provides understanding

Community

Genome Figure 56.1  Diagrammatic illustration of the linkage between ecology and toxicology in ecotoxicology and ecotoxicological risk assessment (adapted from van den Brink, 2008).

treatment to control termites in a structure may be assessed against the benefit of the structure lasting for a longer time and not having to be replaced. Similarly, an agriculturalist may consider the cost of the pesticide used to control an infestation of insects in a fruit crop against the benefit resulting from increased value of the crop to the ultimate consumer, be this in improved quality, better storage properties, or greater value. These types of risk assessment are familiar to most of the public and have been used for thousands of years in the agricultural decision making. They are relatively easy to conduct as the risks (loss of the crop, etc.) are measured in financial units and the cost of the control measures are measured in the same units and simple arithmetic can be used to determine if the risk of use (cost) is worth the benefit (profit). In addition to the use of chemical pesticides, there are several other methods of pest control that are often used alleviate a pest problem. Integrated pest management (IPM), is commonly practiced in agriculture and other pest management situations. IPM involves the use of many procedures for pest control, from cultural practices to biological agent, but all involve risk management decisions similar to the use of a pesticide, although the risks are not as easily measured and the benefits are more indirect. For example, the use of an insecticide may be recommended for the control of an outbreak of an insect pest (A), but it may affect a beneficial organism, such as a parasitic wasp, that would control another pest (B). If a pesticide were used, pest A would be controlled but pest B would be released from natural controls and may increase in numbers to the point that it may damage the crop to a greater extent than pest A. Although this risk assessment considers the effect of the pesticide on the target as well as the non-target organism in the agricultural unit, the measures used to compare risk to benefits can ultimately be expressed in financial units, reducing the decision to more familiar terms. Risk–benefit decisions made during pest control or pest management operations are internalized to the act of agriculture or forest production (Figure 56.2). These decisions are made with very specific uses in mind. The history of the crop or

Figure 56.2  Illustration of risk assessment and risk management decisions in pesticide use in crop production.

animal production system is known and the particular pest situation may be known in great detail, especially in cases where the size of the pest population and information on its life table are available, such as in intensive IPM practices where scouting of pest populations and descriptive information on climate and other environmental factors are collected. While a large number of risk assessment and risk management decisions may be taken as part of the use of pesticides in pest management, these decisions are internalized to the production unit, the agricultural field, the structure, or the health district. Seldom do these decisions consider the effects of the pesticide outside the area of specific use, such as may result from movement away from the area of application to areas that support non-target organisms or to non-target organisms that utilize the agroecosystem as habitat (birds, mammals, or other terrestrial organisms). These external risks are those that are focused on environmental regulations and those that will be dealt with in more detail in this chapter. This focus does not imply that risk assessment in relation to management of pests in a particular situation of crop production or protection of health or property is unscientific or oversimplified; it is merely recognition that risk assessment inside the agroecosystem is different from that outside the agroecosystem. This chapter is therefore devoted to the risk assessment of pesticides as it applies to pesticides that have moved off the agroecosystem and their effect are not part of the risk assessments conducted as part of production or to organisms that are exposed to pesticides because they make use of the agroecosystem as habitat or as a source of food (Figure 56.2).

Chapter | 56  Ecotoxicological Risk Assessment of Pesticides in the Environment

56.1.2  Assessing Risks from Pesticides   in Relation to Other Substances Substances that have adverse effects in the environment have certain combinations of characteristics that lead to the possibility of adverse effects. These characteristics include the inherent toxicity of the substance and the potential for exposure to the substance. The toxicological properties of a substance are determined by its physical and chemical properties and the biochemistry and physiology of the organisms exposed to it. Exposure to the substance is also dependent on interactions between the chemical and physical properties of the substance and its environment. This is as true for pesticides as it is for all other substances. Contrary to public perceptions, pesticides do not possess special properties that make them chemically unique as a class or give them special toxicological or environmental properties that are not found in other substances. Depending on the structure of the molecule, the physical and chemical properties of pesticides (and other substances) span a large range of values. As for other substances, some pesticides are highly persistent, highly mobile and highly toxic. It must be recognized that relatively few substances possess the necessary properties to place them in this category; however, pesticides are usually toxic to at least one class of organisms (otherwise they would not be used as such). However, the physical, chemical, and environmental properties of pesticides span the entire range observed in other substances. The process of pesticide risk assessment allows the proper identification and categorization of these substances according to their risks to the environment.

56.1.2.1  The Need for Pesticide Risk Assessment Pesticide risk assessment plays a crucial role in strategic planning and priority setting and in helping society to determine environmental or other priorities. Risk assessment is used in a number of forms by pesticide regulatory agencies in many countries as well as internationally through the UN’s Food and Agricultural Organization (FAO) and the World Health Organization (WHO), which itself conducts risk assessment because of the use of these substances in the protection of public health. In the United States, the Federal Insecticide, Fungicide and Rodenticide Control Act (FIFRA) specifically requires risk assessment as well as risk–benefit analysis. Canada has a Pest Control Products Act and many other jurisdictions such as members of the European Union, Japan, Australia, etc., also have judicial instruments that require the use of pesticide risk assessment procedures in some form or another.

56.1.2.2  The Basic Concepts of Risk Assessment in Relation to Pesticides The basic concepts of the process of risk assessment have been discussed in a number of documents (SETAC,

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1994; Environment Canada, 1997; EUFRAM, 2006a) and books (Reinert et al., 1998 Suter et al., 2007). The EPA Framework for Risk Assessment (U.S. EPA, 1992, 1998) is a useful general model as many of the other risk assessment schemes are essentially similar. It is important to recognize that, although all risk assessment methods are essentially similar, they may be used for different purposes. In some situations, risk assessment is used to set environmental guidelines and criteria while in others it is used to assess risks in situations where exposures are known and their significance is being assessed. In the first use, the objective is to determine a single protective criterion while, in the second, the task is to determine which of one or more measured or estimated exposures to a potentially hazardous substance have caused unacceptable risks (Figure 56.3). The first step in the risk assessment framework (U.S. EPA, 1992, 1998; EUFRAM, 2006a) is problem formulation or the definition of the problem (Figure 56.3). This is an important step as it lays down the foundation upon which the rest of the assessment depends (Suter et al., 2007). This step defines the objectives and scope of the entire process. This is important, since ecological risk assessments are often complex, involving several biological components ranging from organism through population, community to the ecosystem, and possibly several stressors and/or responses as well. A detailed plan (conceptual model) is required at the onset of the assessment to identify and prioritize all of the issues that need to be addressed as well as the data requirements of the assessment. Communication between the risk assessor and risk manager should occur during this phase as well, ensuring that all information required by the manager is provided by the assessment and that all ecologically relevant issues are also addressed during the assessment. This process may be informal in its initial stages but will become more formal as one or more iterations are made through the risk manager and the risk assessor. The EPA framework for risk assessment (U.S. EPA, 1992, 1998) further divides problem formulation into five subsections: stressor characteristics, ecosystems potentially at risk, ecological effects, endpoint selection, and the conceptual model.

56. 1.2.3  Characterization of the Stressor During characterization of the stressor, all potential stressors are identified and characterized by type (i.e., chemical or physical stressor), intensity, duration, scale, frequency, and timing (whether or not the stress occurs during an important biological cycle). For pesticides, the type of stressor is normally an organic or inorganic substance. Intensity refers to the exposure concentration of the pesticide as experienced by organisms in the relevant environmental matrix (water, soil, etc.). The concentration of the pesticide likely to result from its release into various

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Figure 56.3  Generic framework for developing environmental criteria and for ecological risk assessment (adapted from U.S. EPA, 1998).

matrices may also be useful for identifying organisms most likely to be affected. This is important in assessing risks from pesticides because environmental fate is an important determinant of exposure concentration. Duration and frequency of the release of the stressor are also important. Risk assessments based on long-term, continuous releases will be very different from those based on sporadic, short-term releases. Likewise, labile substances must be assessed differently from more persistent substances. Timing in relation to biological cycles may also be important. This is particularly relevant in noncontinuous releases or the use of stressors such as pesticides where application is usually linked to the biological cycle of the pest organism and, by default, to those of other organisms in the environment. In the process of risk assessment as applied to IPM, concurrent cycles of beneficial organisms are taken into account in the choice of pesticide and the timing of application. In the case of long-term or frequent exposures, organisms may be more sensitive at some times of the year than others (spawning in fish) and risks may need to be assessed in relation to these events. Scale is important, both in terms of the range over which these effects occur and the degree of heterogeneity in the distribution of the stressor. Lack of mixing of an effluent stream in a river or lake may result in differential exposures and the treatment of relatively small blocks of land with pesticides in agriculture and forestry may leave large untreated areas as refugia from which repopulation may rapidly occur. In fact, for many pesticides, annual or more frequent applications are required as pest populations will recover, either from protected individuals or as a result of immigration from untreated refugia or metapopulations. For pesticides, hazard identification is easier than for many other environmental risk situations where the identity of the substance(s) may be unknown. In the case of

pesticides, the identity of the pesticide(s) plus any contaminants or formulants is known (see other chapters on individual pesticides in this book). In addition, environmental breakdown products will be known, as will pathways of metabolism in a number of species. Moreover, because pesticides carry directions for use, the maximum amounts used per unit area are known and so is the frequency of use and region/crop type where the pesticide is to be used. This makes risk assessment for pesticides easier in one sense but may misdirect attention to pesticides because there is generally more information available for them.

56.1.2.4  Characterization of the Receptor Organisms The receptor organisms in the system may be identified or unknown. In the former case, the motivation for the risk assessment may have been the observation of changes in populations or in community structure such as bird or fish kills or changes in distribution of plants. Where the pesticide has not yet been used in the environment, such field observations are not possible and laboratory tests on a range of organisms or systems (cosms; see below) may be necessary to identify populations or communities at risk. These data can then be used to set environmental guidelines or criteria (Figure 56.3) and is often the case when a new pesticide is brought forward for initial registration. However, for pesticides, a great deal of additional relevant information is often available. For example, the mechanism of action may be known and, from this, likely sensitive organisms can be deduced. Even where the mechanism of action is not known, the wide-ranging screening tests used in the development of a pesticide can be utilized to identify most sensitive classes of organisms. The organisms or communities most at risk may also be more easily

Chapter | 56  Ecotoxicological Risk Assessment of Pesticides in the Environment

identified when the use pattern of the pesticide is available to characterize the receiving ecosystem.

56.1.2.5  Characterization of the System A similar process to the above occurs with characterization of the ecosystem potentially at risk. This involves the identification of the ecosystem and the characterization of some of its properties, such as the abiotic factors that may affect the fate of the pesticide and the structure and function of the ecosystem. For example, knowledge of the crops on which the pesticide will be applied and when it is likely to be used will help to identify receiving environments and likely non-target organisms.

56.1.3  Protection Goals, Assessment Endpoints, and Measures of Effect While many jurisdictions have regulatory policy goals such as environmental protection, these are not useful in criteria setting and risk assessment as they are usually ambiguous or difficult to define and measure. Thus, one of the key steps in the problem formulation is the statement

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of the assessment endpoints or what is to be protected. Protection goals are the societal drivers for setting risk assessment endpoints and many of these are linked to sustainability of ecosystems (Calow, 1998; Brock and Ratte, 2002). In protecting the environment, the intention is that populations and communities be sustained in the environment and that this involves protection from change resulting from a manageable source of risk, not just decreases. Adverse responses in the ecosystem are usually misperceived as always being negative; a decrease in population or a decrease in a function. Increases in populations, such as algal blooms, or functional processes, such as oxygen consumption, may be just as deleterious in the ecosystem (Giesy, 2001). To maintain ecosystems in a pristine condition may be important for certain environmental uses, such as nature reserves. For agroecosystems and their important benefits to humans, this is not practical. There are three general categories of protection goals related to the sustainability of communities and ecological functions and services. These relate to ecosystem structure, function, and aesthetic value to humans (Table 56.1). The structure of an ecosystem is a combination of which organisms are present and how

Table 56.1  Examples of Ecosystem Protection Goals, Assessment Endpoints, Measures of Response, and Toxicological Effect Measures for Use in ERA Protection goal No unacceptable change in overall species richness and density, population densities of key or important ecological species, or population densities of indicator species

Assessment endpoint

Observed response

Toxicological measure of effect

Probability of a 5% reduction in diversity

Toxicity of the pesticide to representative species tested in the laboratory or field

Acute or chronic LC/LD50s or NOECs in several species of aquatic or terrestrial organisms

Probability of a 10% reduction in number of songbirds

Toxicity of the pesticide to birds tested in the laboratory lethality, reproduction, development

Acute or chronic LC/LD50s or NOECs for test species such as canaries, quail, or mallard ducks

Probability of a 1% loss of individuals of a population of endangered species

Toxicity of the pesticide to a surrogate organism for the endangered species

Acute or chronic LC/LD50s or NOECs for the test species

Protection of biogeochemical cycles and energy flow

Probability of a change in ecosystem functioning and functionality by 10%

Effect of the pesticide on a process such as photosynthesis or microbiological transformation (nitrification)

Acute or chronic EC50, EC10, or NOEC for photosynthesis or nitrification

Perceived aesthetic value or appearance of the ecosystem/ landscape

Probability of loss of species with a popular appeal but not essential in the ecosystem

Toxicity of the pesticide to representative species tested in the laboratory or field

Acute or chronic LC/LD50s or NOECs for the test species

Probability of 5% increase in algae in lakes

Toxicity of the pesticides to herbivores that would normally consume the algae

Acute or chronic LC/LD50s or NOECs in several species of aquatic herbivores

Modified From Brock and Ratte, 2002; Solomon et al., 2008a.

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many there are, while function relates to the role of these organisms in ecosystem processes. The choice of protection goals (and by extension, assessment endpoints) may be based on ecological knowledge or on human value judgments. For example, there is a propensity to select functional protection goals and assessment endpoints when the populations of the potentially affected organisms may change rapidly for natural reasons, recover from effects rapidly, or are difficult to characterize. Examples are bacteria and fungi in soil and sediment or algae in aquatic systems. For populations that have lower recovery potential, are easily characterized and/or greatly valued, structural protection goals, such as absolute population numbers, are more generally used. Examples of these are fish or birds. Choices of protection goals may also be determined on the basis of value judgments, for example, if the activity that generates the risk also brings great benefits, structural changes may be tolerated if functions are unaffected. These three ecological response categories may be further subdivided, as shown in Table 56.1. Assessment endpoints must be carefully chosen and must be unambiguous (not policy goals). Poorly selected assessment endpoints have resulted in more risk assessment failures than any other possible error in risk assessment (Suter et al., 2007). Assessment endpoints are explicit expressions of the actual values that are to be protected, are the ultimate focus in risk assessment, and act as a link to the risk management process, such as the policy goals. Assessment endpoints usually have the following characteristics: They should be expressed in a way that they can be quantified (this allows achievement to be gauged), be ecologically relevant, be susceptible to the stressor, and should have societal value. This last criterion, social value, must be treated with caution as it is subject to perceptual interpretation. For example, societal values may change over time. An example of this is the public regard for wetlands (Suter et al., 1993). Before the 1970s, the public saw little value in wetlands and many were, in fact, drained to reduce nuisance from mosquitoes or to increase cropland area. Since the public now better understands the linkage between wetlands and amphibians, birds, waterfowl, and flood protection, the social value of wetlands has increased and their functions are now accepted as worthy of protection. This suggests that we should seek endpoints that have ecological as well as societal relevance. This may require that the risk assessor educate the public on the value of the ecologically relevant endpoints. Measures of effect are exposures that cause responses, usually in laboratory studies, but are related quantitatively or qualitatively to the assessment endpoint. Therefore, measures of effect should be based on responses that are biologically significant. For example, a physiological or biochemical change in an organism (often referred to as a biomarker or bioindicator) must be related to the survival or fecundity of the population before it can become a useful measure of

Hayes’ Handbook of Pesticide Toxicology

effect. The bioindicator is useless unless it is linked to relevant effects at the population level in organisms that are ecologically important in the ecosystem. A number of factors must be taken into account in selecting measures of effect (U.S. EPA, 1998). These include consideration of indirect effects, sensitivity and response time, diagnostic ability, costeffectiveness, and ease of measurement. When an assessment endpoint can be measured directly, the assessment endpoint and measures of effect are the same. However, this is seldom the case. Examples of assessment endpoints and measures of effect are shown in Table 56.1.

56.1.3.1  Function in Ecosystems and Ecological Risk Assessment of Pesticides Assessment endpoints can be viewed from the position of “prevention of pollution,” “ecological thresholds,” “recovery,” or “redundancy” (Solomon et al., 2008a). Although assessment endpoints and measures of effect can be defined at all levels oforganization in ecosystems, from the individual to the community (Figure 56.1), these are not necessarily of equal importance (Suter et al., 2007). When conducting risk assessment of pesticides in the environment, individual organisms in the ecosystem are generally regarded as transitory and, because they are usually part of a food chain, are viewed as individually expendable (Suter et al., 2007). A self-maintaining or reproducing population is persistent on a human time scale and can be easily conceptualized by humans as being in need of protection. Thus, most assessment endpoints in ecological risk assessment are defined at the level of the population or the functions of populations, communities, and ecosystems, rather than at the individual organism level. This acknowledges the fact that populations are less sensitive than their most sensitive member and, likewise, that communities and ecosystems are less sensitive than their most sensitive components. Effects on a population are not necessarily of concern (to the ecosystem) as long as the functions the population can be taken over by other organisms. In this context, function is the interaction of the population with other populations or the abiotic environment. Functions in ecosystems are normally related to energy and nutrient flow: production of biomass (primary production), consumption of biomass (grazing or predation), controlling the abundance of other (prey) species, providing food to predators, or processing organic detritus, such as shredding plant tissue, macerating animal remains, and mineralizing organic compounds (Suter et al., 2007). Resiliency, functional redundancy, and recovery from adverse effects are essential to the continuance of ecosystems in the face of natural stressors. Resiliency is the ability of individual organisms or populations of organisms to tolerate some stress and usually involves changes in biochemical and/or physiological processes. Redundancy is where multiple species are able to perform each critical function

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Figure 56.5  Illustration of extrapolations within and between systems in the environment (adapted from van den Brink, 2008).

Figure 56.4  Illustration of ecosystem resiliency and redundancy in response to stressors and effects caused by interactions between organisms (adapted from Solomon and Takacs, 2002).

(Walker, 1992; Baskin, 1994; Walker, 1995) and is the result of selection imposed by variable and unpredictable environmental conditions. Most ecosystems exhibit functional redundancy and this is particularly relevant to the ecotoxicological risk assessment of pesticides. It is the basis for being able to tolerate effects in some sensitive populations as these are unlikely to impair the functions of the ecosystem as a whole, such as in setting water quality guidelines (Stephan et al., 1985). As is illustrated in Figure 56.4, there is a general relationship between exposure concentration and impact of any substance; however, there are deviations from this general rule. For example, functions of ecosystems may be maintained where few species are affected but, as the number of species affected increases, indirect effects amplify the effects of the substance to greater than predicted levels. Redundancy of function has been observed in a number of experimentally manipulated systems ranging from terrestrial (Tilman, 1996; Tilman et al., 1996) to aquatic (Stephenson et al., 1986; Giddings et al., 1996; Giesy et al., 1999; Giddings et al., 2000, 2001, 2005). These observations support the concept that, in ecotoxicological risk assessment, some effects at the level of the organism and population can be allowed, provided that these effects are restricted on the spatial and temporal scale. In other words, they do not affect all communities all of the time and that keystone organisms are not adversely affected. In the context of selecting assessment endpoints, it has become increasingly recognized that these should be at the functional level of populations and the community and that some effects on populations and species diversity may thus be tolerated. It is possible that a chain of events can occur whereby effects on one population may cause ripple effects throughout the ecosystem. These effects would occur in keystone species. Keystone species often supply physical habitat or modify the habitat in a way that cannot be replicated. Thus, removal of habitat (other populations) may be the

root cause of the risk to a population designated for protection. An example of this is seen in the case of the spotted owl in the Pacific Northwest and removal of nesting habitat (James, 1994). Effects of pulp mill effluents on larval and juvenile fish in the Baltic is another example of the importance of habitat (Lehtinen et al., 1991), as is the relationship between sea urchins, kelp, and sea otters (Estes et al., 1998). Examples of these types of keystone responses are infrequently reported for pesticides in the nonagricultural environment but are a very real problem in IPM situations within the agroecosystem.

56.1.3.2  Extrapolation in Ecotoxicological   Risk Assessment of Pesticides Extrapolation (Figure 56.5) is the way to understand consequences, to provide explanation, and to link measures of effect to assessment endpoints and is very important in all risk assessment procedures (Solomon et al., 2008a; van den Brink, 2008), and is discussed in greater detail in a recent book (Solomon et al., 2008b). The costs of obtaining measurements of endpoints at the population level are often very great and, as a result, the bulk of toxicological testing has been focused at the organism level. It is often necessary to extrapolate from data at the level of the organism to the level of the population or community. This concept is not foreign to risk assessment as most human health risk assessments are based on interspecies extrapolation from rodents and other test animals to humans. Use of data from individual organisms in the characterization of risk to populations is based on the belief (or dogma) that protection of individuals will protect the population and that this will also protect the structure of the community, and, ultimately, the functioning of the ecosystem. A number of types of extrapolations are routinely used in risk assessments. These can be categorized on the basis of the starting point, desired outcomes, and factors that can influence and/or confound the extrapolation. These have been broadly divided into two types, range and data extra­ polations, and are summarized in Table 56.2.

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Table 56.2  Summary of Types of Extrapolation Used in Ecotoxicological Risk Assessment of Pesticides Type

Explanation

Influenced by Range extrapolation

Extrapolation beyond the observed data but within the same situation of species and substance.

Responses are inferred or estimated outside the range of the observed data from which the relationship was derived.

The model chosen to describe the relationship and the model for the extrapolation. Biological thresholds and low-exposure hormesis may affect the extrapolation.

Matrix/media extrapolation Extrapolation between matrix and/or media.

Responses in one medium or matrix are inferred from those observed in another, such as between freshwater and saltwater, sediments of different organic matter composition, waters of different hardness.

Chemical and physical interactions between the toxicant and components of the medium. Complexation of metals or adsorption of organic substances can influence exposure of the organism. Can also be mediated by interactions between the matrix and the physiology and biochemistry of the organism such as in competition for transport sites.

Temporal extrapolation Extrapolation between exposure regimes.

Responses to chronic exposures are inferred from those observed in acute exposures, pulsed from continuous exposures, one route of exposure from another (oral vs. matrix, dermal vs. matrix, etc.) Because of kinetics, body burden may be dependent on exposure time.

Toxicokinetics is the primary determinant of differences in body burden, target tissue concentration, and responses to different exposures. An issue is the nature of the response, which may differ for acute to chronic testing. In this extrapolation, an assumption is made that potency under one condition of exposure is proportional to that under the other conditions.

Data extrapolation Extrapolation across age and/or developmental stage.

Responses in one life-stage are inferred from those in another, such as from juvenile to adult or vice-versa or from one sex to another. Is often subsumed by the use of the effect data for the most sensitive life-stage.

The assumption that juvenile stages are more sensitive is not necessarily always true. Cyclical activity such as reproduction and molting may make organisms more sensitive at certain seasons or stages of the life cycle and this may be confined to certain classes of substances such as the endocrine modulators.

Extrapolation between species.

Responses in one species are inferred from those in another. Often subsumed through the use of effect data from the most sensitive surrogates used in laboratory studies to those organisms that occur in the environment.

Normally by uncertainty factors which vary according to the measures of effect and the value of the organisms being protected. The method is generally believed to be protective of responses in the field because of the use of uncertainty factors and the maximal exposures that occur in most laboratory tests.

Extrapolation between levels of biological organization.

Responses in one system are inferred from those in other systems. Extrapolation from bioindicator at the cellular or physiological level to organisms, from organisms to populations, from populations to ecosystems.

Well calibrated bioindicators can be used as predictors of effects at the population level but they may lack specificity or not respond in a consistent manner to the stressor. Extrapolation from the organism to the population is usually by way of models. These may be affected by uncontrolled or unknown environmental factors or by incorrect parameterization.

Extrapolation from species to communities.

Responses in communities are inferred from responses in tests conducted with many species, such as in the use of species sensitivity distributions (SSDs) of effect measures to extrapolate laboratory data to communities.

Size of the data set and types of organism tested can influence the representativeness of the laboratory data and the model used to characterize the data. Incorrect combinations of species may confound extrapolation of group-specific stressors such as pesticides.

Extrapolation from one location to another.

Responses in one geographic zone are extrapolated to another, i.e., temperate to tropical.

Differences between species composition between the locations and in the effect of local conditions on the fate of the substance.

Extrapolation from one substance to another.

Responses to one substance are inferred from responses measured in another, which shares the same mechanism of action.

Available data from field and/or laboratory studies must be sufficient to reduce uncertainty. Most applicable to well-studied substances, such as pesticides, where the mode of action is understood.

Adapted from Solomon et al., 2008a.

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56.1.3.3  Uncertainty in Risk Assessment   for Pesticides Uncertainty is a very important component of risk assessment because it influences the probability (risk) that an adverse effect will occur. As has been pointed out (Suter et al., 2007), it is necessary to understand the uncertainties associated with the data that are used to make risk management decisions. Traditionally, risk assessors have used worst-case conservative assumptions of exposure and response as well as uncertainty factors which, depending on the amount of data, may range from 1 to 10,000 (Zeeman and Gilford, 1993; Environment Canada, 1997; European Union, 1997; OECD, 2002) rather than actually estimating uncertainty. This approach has several drawbacks because worst-cases scenarios ignore the probability of occurrence, may not be multiplicative or additive, are inconsistent as it is always possible to conceive of a still worst case, and are based on the premise that there are no societal or environmental costs resulting from the regulation of false positives (Suter et al., 2007). Use of worst-case assumptions is only applicable in lower tiers of risk assessment where data sets are small and there is greater uncertainty.

56.1.3.4  Use of Conceptual Models in Risk Assessment of Pesticides The conceptual model is an aid in the development of working hypotheses as to how the pesticide might affect components of the ecosystem. The conceptual model includes a description of the ecosystem potentially at risk and the relationship between assessment endpoints and measures of

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effect and helps to narrow the questions and hypotheses to a manageable number. In the conceptual model, a preliminary analysis of the ecosystem likely to be impacted, characteristics of the stressor, and likely ecological effects is used to define possible exposure scenarios for receptor(s). Exposure scenarios consist of a qualitative description of how the various receptor organisms co-occur or may come into contact with the stressor. These descriptions consider a number of issues to better match the conceptual model to the actual field situation. These include, inter alia, the medium; the geographic, spatial, temporal and biological scales; and the amount of detail needed. Although many hypotheses will likely be formulated during the problem formulation phase, only those that are considered most likely to contribute to risk should be selected for further evaluation in the analysis phase. Constructing the conceptual model is not necessarily a single-pass process. Several iterations may be necessary before a sufficient level of understanding has been attained to proceed with the final model and the analysis itself. A generic conceptual model for exposure pathways for pesticides in aquatic systems is illustrated in Figure 56.6. Not all of the exposure pathways will be applicable in every situation and some may be more important than others. Similar models may be constructed for terrestrial environments and also for effects. By refining these models, a potentially very complex set of risk hypo­theses can be reduced to a more manageable size.

56.2  The risk assessment analysis After the risk assessment problem and conceptual model have been defined, the next steps in preparing for the

Figure 56.6  Generic conceptual model for exposure pathways for pesticides in aquatic systems. The arrows show possible exposure routes for direct effects. Indirect effects are not shown.

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ecological risk assessment are the characterization of the effects of the pesticide and exposures to the pesticide.

56.2.1  Characterizing Effects For organisms exposed through the matrix they inhabit (soil, air, or water), measurement of effects involves the consideration of three variables: exposure concentration, exposure duration, and response. Exposure concentration is a continuous variable and the proportion responding is a quantal variable. Duration of exposure is usually a continuous variable but, under field conditions, may be complicated by temporal variations in the concentration. In the simplest case, the proportion of organisms responding is a function of concentration. This relationship is nonlinear but may be analyzed by a linear model where the concentration is expressed as a logarithm and the percentage of organisms responding as a probability (probit scale). The most commonly reported number with respect to toxicity is the median response concentration. This the concentration at which half the organisms respond with the defined response. Generally, the defined response is lethality (LC50) but it may also be a specific effect such as immobility of small organisms (EC50). The general use of the 50% response is because of the higher precision of the estimate and the narrower confidence intervals around this point. In practice, risk assessors are usually more interested in low or high proportions of responses. Other responses that a commonly used are LOAEC (lowest observed adverse effect concentration), the lowest concentration at which adverse effects are observed (this depends on the range of doses tested) or the NOAEC (no observed adverse effect concentration), the highest concentration at which no effects are observed. Because the NOAEC and LOAEC are dependent on the choice of concentrations used in the study, there is a move toward using a low level of response derived from a regression of response and exposure (Crane and Newman, 2000). This is similar to the derivation of the benchmark response suggested for use in the interpretation of mammalian studies (U.S. EPA, 1995b). It is important to recognize that time is a critical parameter in the relationship between exposure and effect. This is particularly the case when exposure is through the matrix and uptake kinetics affect the concentration of the substance in the body of the organism. It is also important when pulsed exposures are considered. In almost all cases in the environment, exposures to pesticides are not to constant concentrations but to some form of pulse, the peak and duration of which is determined by the interaction between the pesticide and its environment. After initial application of the pesticides to the matrix, breakdown, adsorption, or hydraulic dilution (in aquatic systems) will occur. Exposures may thus be of unpredictable duration unless the specific characteristics of the receiving system

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are known. In practice, they range from very short to occasionally very long. Few compounds show a linear relationship between exposure concentration, exposure time, and effect (Giesy and Graney, 1989). For instance, pulsed exposures to greater concentrations may be less toxic than longer-term exposures to lesser concentrations. Although these exposures result in the same integrated concentration  time exposure (g/L  hours), they may not account for pharmacokinetic or repair processes, which control the effects that are observed. In answer to this concern, it has been suggested that customized toxicity tests of a specific duration be carried out (SETAC, 1994). A similar result can be achieved by assessing mortality at earlier time periods (e.g., 4 h, 8 h, etc.) in a 48- or 96-h study, which would allow data to be created that compares short-term exposure to short-term toxicity (ECOFRAM, 1999a). This has the additional advantage of improving the statistical analysis to allow estimates of the effects (LC50s or EC50s) to be made more accurately (because more observations are available) and allows the estimation of responses for any time period between those at which measurements have been made (ECOFRAM, 1999a). However, it should be noted that observation of responses at a particular time interval during a test does not necessarily mean that toxicity will not change after that time if the organisms are moved to an uncontaminated medium (latency of response). Very few substances are instantly toxic and some time may have to elapse before the organism shows maximal response even though exposure has ceased.

56.2.1.1  Measuring Effects of Pesticides   in Individual Species Standardized test methods for pesticides are routinely used and required by a number of regulatory agencies for the registration of pesticides (CFR, 2004; OECD, 2008). As test methods for pesticides are constantly being updated, changed, and added to, only the general principles of the tests will be summarized. The basic principle behind the use of standardized laboratory toxicity tests is not that the particular organisms used in the test are those that require protection in the environment, but rather that these organisms act as surrogates for all those other organisms in the ecosystem that could be exposed but which, for one reason or another, cannot be tested in the laboratory. Because of this, test organisms are usually selected for ease of use (easy culturing techniques or easy availability in the environment) and because historical testing has shown that the species is particularly sensitive and would therefore provide a worst-case measure of effect. To make the effect measure even more conservative, the tests are normally conducted under conditions where the exposures are maintained at a constant concentration, usually by continuous addition to a continuous flow treatment

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system. Most regulatory organizations now require that all toxicity testing be conducted under Good Laboratory Guidelines (GLP) to ensure that the tests have been carried out appropriately and that proper quality controls have been implemented (OECD, 1998; U.S. EPA, 2007b). (a) Tests with Terrestrial Organisms For terrestrial systems, plants, microorganisms, inverteb­ rates, and mammals may be tested. Testing of terrestrial plants is not yet widely required for pesticide risk assessment but it should be recognized that a wide range of plants, including both crop and weed plants are routinely tested in the discovery and development of a pesticide. These data give a general indication of sensitivity within classes of plants and can be used in assessing likely sensitive non-target species. Endpoints used in plant tests include germination, wet or dry biomass, or root elongation. For soil microorganisms (bacteria and fungi), functional responses, such as nitrification (aerobic), denitrification (anaerobic), heterotrophic nitrification, and mineralization of nutrients may be used to derive measures of effect. Tests on terrestrial invertebrates are directed toward assessing effects in beneficial invertebrates such as the honey bee, Apis mellifera or other beneficial arthropods, various types of earthworms, such as Eisenia fetida, Lumbricus terrestris, or Enchytraeus albidus, and Colembola (springtails), such as Folsomia candida. Tests of terrestrial vertebrates are normally focused on mammals and birds. For mammals, the extensive testing on rodents and other laboratory test species used in the process of human health risk assessment (see Chapters 10 and 14) is normally used as a surrogate for mammalian wildlife. Birds used in toxicity tests include the bobwhite quail, Coturnix coturnix; Japanese quail, Coturnix coturnix japonica; leghorn cockerel, Gallus gallus; mallard duck, Anas platyrhynchos; and ring-necked pheasant, Phasianus colchicus. In most cases, oral toxicity is measured either as a dose or as a concentration in the diet. Long-term feeding studies for either chronic or lifetime exposures may also be undertaken. (b) Tests with Aquatic Organisms For aquatic systems, plants, invertebrates and vertebrates are tested. Tests on aquatic plants are required for most herbicide registrations with the U.S. EPA (CFR, 2004) and in Europe (European Commission, 1994). Organisms used include algae such as the freshwater algae, Pseudokirchneriella subcapitata, Anabaena flos-aquae, Microcystis aeriginosa, and Navicula peliculosa and the saltwater algae, Chlorella spp., Chlorococcum spp., Dunaliella tertiolecta, Isochrysis galbana, Nitzchia closterium. Skeletonema costatum, and Porphyridium cruentum. The freshwater macrophyte Lemna spp. (duckweed) is also used. Measures of effect include growth (as a percent of control) and cell numbers are usually reported as

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EC50s or IC50s. Measures of effect for algae are based on growth and the tests are normally run for periods of time that include many generations. Thus the EC50s reported are similar to those that would be expected from chronic studies in other organisms. For invertebrates in freshwater, the water fleas, Daphnia magna and Ceriodaphnia spp., are used for acute assay and chronic studies that include one life cycle. The midge, Chironomus spp., is normally used for acute assays only. For saltwater invertebrates, shrimp, such as Mysidopsis bahia and Penaeus duorarum, are used for acute and chronic assays for reproduction, mortality, and growth. The oyster embryo-larval test Crassostrea spp. is also used in an acute assay for effects on mollusca. The need for testing saltwater species is usually dictated by the use of the pesticide and the potential for contamination of saltwater systems such as estuaries. For aquatic vertebrates, standardized tests include acute life-cycle tests and early life-stage tests. Some assays have been reported with amphibians (mostly larval stages) and are likely to become more widely required with increased interest in amphibian declines. Freshwater species used in acute tests include the rainbow trout, Onchorynchus mykiss; brook trout, Salvelinus fontinalis; channel catfish, Ictalurus punctatus; fathead minnow, Pimephales promelas; and bluegill, Lepomis macrochirus; while saltwater fish include the sheepshead minnow, Cyprinodon variegates; mummichog, Fundulus heteroclitus; longnose killifish, Fundulus similis; silverside, Menidia spp.; and the threespine stickleback, Gasterosteus aculeatus. Early life-stage tests may also be conducted. These tests focus on development and growth during the first 30–90 days of development from the egg. Species used include the fathead minnow, bluegill, brook trout, flagfish, Jordanella floridae; and sheepshead minnow. Depending on several factors related to fate of the pesticide and results observed other tests, full life-cycle tests may be required. Test organisms include fathead minnow, bluegill, brook trout, flagfish, and sheepshead minnow. Measures of effect for acute studies normally include mortality, while early life-stage and life-cycle tests include measures related to growth, development, and reproduction. Other tests for responses such as endocrine disruption are under development (Ankley et al., 2001; U.S. EPA, 2007a). (c) Sources of Data and Data Quality Although the above organisms are favored for regulatory testing, data from other species are also considered in pesticide risk assessment and may be submitted to the registration authorities, obtained from the open scientific literature, or databases such as ECOTOX (U.S. EPA, 2001a). In these situations, the studies may not have been conducted according to specific guidelines or under goodlaboratory practice protocols (OECD, 1998) and the data must be used with appropriate critical review. For aquatic

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organisms, the Quality Standards for the Great Lakes System (GLI U.S. EPA, 1995a) or those of CCME (2007) for acceptance of a study are useful. Preferred studies are those conducted according to guidelines, under flowthrough exposure conditions, and with the concentration of the pesticide measured. (d) Types of Organisms Used in Toxicity Tests Many substances such as pesticides have some degree of specificity in their mechanism of action and it is best to separately consider risks to different groups of organisms. For example, an insecticide that acts on the nervous system of insects is unlikely to be highly toxic to plants; however, organisms with well-developed nervous systems (insects, other arthropods, and vertebrates) are likely to be more sensitive. Therefore, it makes sense to separate plants and animals in the assessment process as they will respond differently. Specificity of action may not always be the case. For example, some biocides, which target a general target such as energy production (chlorophenols) are similarly toxic to a wide range of organisms (Liber et al., 1994) and grouping all organisms together for risk assessment may be appropriate. Thus, from a basic understanding of the mechanism of action of a pesticide, it may be possible to identify and group organisms by sensitivity to the pesticide. This is helpful from the point of view of risk assessment because it allows the assessor to focus on the groups at greater risk and to devote less time and resources to groups that are at small or negligible risk (Campbell et al., 1999; Giddings et al., 2002). In addition, with knowledge of the ecology of the potentially impacted system, it is possible to assess the likelihood that indirect effects will occur as a result of an effect on keystone groups of predator or prey/ food organisms (Campbell et al., 1999; Liess et al., 2005). The habitat of the organisms may also be important. For example, there may be good mechanistic reasons to separate data for freshwater and saltwater organisms where it is known that one group has an inherently different sensitivity because of interactions between salinity and the stressor of concern (Hall and Anderson, 1995). In other cases, this may not be necessary. Some of the differences reported between saltwater and freshwater animals (Wheeler et al., 2002) may be because of a difference in the makeup of the community; there are very few species of insects found in saltwater. There has been much debate about differing sensitivity of species from different geographical locations. Because of historical regulatory requirements in North America and Europe, most toxicity data are available for Nearctic (North American) and Palaearctic (European, in particular, the western Palaearctic) organisms (Brock et al., 2008). Although there are similarities in the fauna and flora of these two regions, there are also differences in biodiversity and community composition. A few studies have compared the relative sensitivity of Nearctic and Palaearctic species.

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Reanalysis of available toxicity data for atrazine (Solomon et al., 1996) and copper (Brix et al., 2001) revealed no difference in the median sensitivity of Palaearctic or Nearc­tic fish species to copper but a greater sensitivity of Nearctic fish to atrazine (Brock et al., 2008). Sensitivity of arctic invertebrates to metals was similar to invertebrates from other climatic regions (Chapman, 1993). When comparing the sensitivity of Nearctic and Palaearctic arthropods to four insecticides (chlorpyrifos, fenitrothion, diazinon, and lindane), Maltby et al. (2005) reported that differences were not statistically different. In the same study, using toxicity values for arthropods, they reported that toxicity values for chlorpyrifos, fenitrothion, and carbofuran in tropical organisms were not statistically significantly different from those of temperate organisms. Others have reported differences between tropical and temperate species of animals for some chemicals (Kwok et al., 2007). The reasons for this may be the temperature of the tests (greater for tropical animals, which may affect toxicokinetics) or that the proportion of taxa comprising data sets were different, for example, fish were more highly represented in most of the data sets from tropical locations. It is also possible to group organisms on the basis of reproductive strategy and life cycle. Thus, organisms that are able to recover rapidly from an adverse effect at the population level (reduction in numbers caused by mortality) may be considered differently from another group of organisms that require a longer period of recovery. For example, most microalgae have short reproductive cycles and would be expected to recover from a decrease in population more rapidly than a population of fish subjected to a similar reduction. Thus, the frequency of occurrence and the intensity of the effect that could be tolerated would be different. This is also important when deciding how the exposure data should be analyzed in terms of frequency of occurrence. The concept of trophospecies (Yodzis and Winemiller, 1999) and trait-based analyses (Baird and van den Brink, 2007) have been suggested for use in characterizing stress ecology and could also be applied to separations of toxicity data for ERA or pesticides.

56.2.1.2  Measuring Effects at the   Ecosystem Level Toxicity studies conducted on a single species in the laboratory cannot take into account effects that involve interactions between populations of different species in communities or those that affect ecosystem function, such as recovery and changes in function. Experimental testing of large populations of humans with a pesticide would be unlikely to receive public approval; however, such studies are possible with ecosystems or subsets of ecosystems. A number of procedures have been proposed for ecosystem- and community-level tests and there are numerous examples of their utility (Hill, 1994; Culp, 2003;

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de Jong, 2008). Most of this work has been carried out in aquatic systems but some terrestrial systems have also been used (van den Brink et al., 2005). The aquatic systems range from simple laboratory systems to complex flowing stream systems, usually referred to a mesocosms or microcosms (cosms). Cosm studies with pesticides provide measures of effect that are closer to the assessment endpoints, for several reasons (Solomon et al., 1996; van den Brink, 2008; van den Brink et al., 2008). Measurements of productivity in cosms incorporate the aggregate responses of many species in each trophic level. Because organisms will likely vary widely in their sensitivity to the stressor, the overall response of the community may be quite different from the responses of individual species as measured in laboratory toxicity tests. Cosm studies allow observation of population and community recovery from the effects of the pesticide and indirect effects of pesticides on other trophic levels. Indirect effects may result from changes in food supply, habitat, or water quality. Cosm studies can be designed to approximate realistic stressor exposure regimes more closely than standard laboratory singlespecies toxicity tests. Most studies, especially those conducted in outdoor systems, incorporate partitioning, degradation, and dissipation, important factors in determining exposure. These factors are rarely accounted for in laboratory toxicity studies, but may greatly influence the magnitude of ecological response (Bernal et al., 2009). One of the major issues in cosm studies is the reporting of the responses that are observed. This has been discussed in workshops (Giddings et al., 2002), and a number of terms are used: NOECPopulation, the greatest concentration tested with no effect, single species; NOECCommunity; greatest concentration tested with no effect, community of populations; NOEACE, no observed ecologically adverse effect concentration; EAC, ecologically acceptable concentration; and RAC regulatory acceptable concentration. Specialized statistical procedures are available to summarize effects in cosms (van den Brink and Ter Braak, 1999), and categories of response have been proposed (Brock et al., 2006). Grouping of species in microcosm by traits or into trophospecies may be helpful in summarizing results in an ecological sense (Wilson et al., 2004; Hillis et al., 2007).

56.2.2  Characterizing Exposure Millions of tons of pesticides move through the global ecosystem each year. During this process, many of these substances may be transformed into other products. Organic compounds are eventually transformed into simple compounds such as carbon dioxide, ammonia, and water (mineralized) or incorporated into the biological cycle via small carbon units such as acetate. However, some may be rendered more toxic or more bioavailable in this process: for example, the conversion of DDT to DDE. Pesticides

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may be transformed, transported, or transferred to another matrix by a number of processes that are controlled by certain environmental parameters (Table 56.3). With the exception of spills, concentrations of pesticides in environmental matrices are small and rates of transformation usually follow zero- or first-order kinetics. For each pesticide, in each matrix, one or two processes usually dominate; however, the relative importance of these may change over time or space (i.e., photolysis may dominate during the day and hydrolysis during the night). Most toxicity tests for environmental risk assessment of pesticides make use of exposures via the matrix in which the organism is present. Therefore, in the characterization of exposures in the environment, the concentration in the matrix is most important. Body dose could be measured in toxicity tests but the equivalent measures from environmental samples may be difficult to obtain.

56.2.2.1  Measuring Exposure Measuring exposure in environmental matrices is one of the critical components of risk assessment but is subject to errors through improper sampling techniques and sometimes by incorrect analyses. Obtaining an unbiased Table 56.3  Environmental Dissipation Processes, Specific Driving Parameters, and Matrix Process

Parameter and matrix Degradation

Photolysis

Intensity of light (water and atmosphere)

Oxidation

Concentration of oxidant (water, air, and soil)

Reduction

Concentration of reductant (water, air, and soil)

Hydrolysis

Temperature, pH (water)

Biotransformation

Presence of populations of organisms, nutrient concentrations, temperature, pH, and other factors that affect the energetics of organisms

Rain-out

Rate of precipitation, sticking coefficient (atmosphere)

Volatilization

Henry’s constant, surface texture (water and soil)

Sorption

Amount of organic matter, lipid, clay content (soil, sediment)

Transport with the matrix

Velocity of wind (atmosphere), current (water), percolation (ground water), transport of particles (soil runoff)

Bioconcentration

Content of lipid in organisms

Movement

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and representative sample from an environmental may be very difficult and costly and yet is probably the most important part of any exposure characterization. Sampling needs to consider both temporal and spatial heterogeneity of the pesticide residues. For example, the concentration of a pesticide may vary with water depth or distance from the shore immediately after a spray-drift contamination of water. Similarly, the concentration of a pesticide in flowing water may decrease over distance from the source of contamination due to breakdown in the water, adsorption to sediments, or dilution from uncontaminated water entering downstream of the source of the pesticide. Concentrations in soil may vary with the crop and soil type and with the chemical and physical properties of the pesticide as well with climatic factors such as rainfall and percolation or leaching through the soil. The objective of sampling the environmental matrix is to obtain a characterization of exposure that will be useful in the risk assessment process. Even with a good sampling design to address spatial heterogeneity, temporal variations in concentrations may be very important in assessing risks in relation to duration of exposure and choosing the appropriate exposure time for the toxicity test data. Because of hydraulic flows in a headwater stream system, peak exposure concentrations in these systems may be very narrow and may be easily missed with a single daily grab sample. They would be incorporated into a continuous sampling system where daily integrated sampling was carried out but very narrow peaks would be obscured. Sampling intervals should be designed to take into account the known hydraulics and breakdown kinetics of the pesticide in question. Thus, in small headwater streams, more frequent sampling with a frequency of less than 1 day may be more appropriate. For slow-flowing rivers, or a rapidly degrading pesticide in a pond or reservoir, daily sampling may be adequate. For slowly degraded pesticides in stagnant pools, ponds, or reservoirs, even less frequent sampling may be needed. The concentration-time series of data that results from this type of sampling can then be analyzed by means of a post-processor tool such as the Risk Assessment Tool to Evaluate Duration and Recovery (RADAR) developed by ECOFRAM (ECOFRAM, 1999a). This tool provides information on pulse height, width, and interpulse interval, which is particularly useful for assessing likely effects on classes of organisms with known recovery times and timeexposure responses (Solomon, 1999). Modern analytical methods may be exquisitely sensitive and be able to detect pesticide residues at amounts as small as 1012 g. In reporting residue data, it is also normal to report the limits of detection and the limits of quantification. The limits of detection (LOD) are method-dependent and are often determined by the specific analytical equipment. In essence, the LOD is that amount that is statistically different from the blank (American Chemical Society Subcommittee on Environmental Analytical Chemistry,

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1980). The limit of quantification (LOQ) is the smallest concentration to which one can assign a definite numerical value with confidence. This concentration is greater than the LOD and is dependent on the repeatability or precision of the analytical method. In practice, it is recommended that a response below the LOD is reported as not detected (ND). A measurement between the LOD and the LOQ can be reported as numbers along with the LOD in brackets or as a trace. Measurements of environmental concentrations of pesticides may contain a significant number of values below the LOD or LOQ. Conventionally, if a mean concentration is to be calculated, those numbers below the LOD/LOQ are assumed to be half the LOD. It is unlikely that all of the NDs would be exactly equal to any single fraction of the LOD. Therefore, a more realistic estimator of these concentrations is to assume a continuation of the distribution of the detected concentrations. Uncensoring methods can be used to estimate these numbers but these would only be useful if a mean were to be calculated. A mean concentration that conceals all of the information inherent in a temporal or spatial array of data is not very useful in ecological risk assessment.

56.2.2.2  Estimating Exposures In many ERAs, the actual concentrations of pesticide in the environment cannot be measured and risk assessors must make use of models to predict what these concentrations will likely be. All models, regardless of complexity, require input data. In general, as the problem definition becomes more complex and highly resolved, the model must become more complex and sophisticated. More complex models generally require more input data and a greater understanding of the system and the input data usually must be available at greater resolution. If the model is too simple, it may not be capable of answering the complicated question. On the other hand, if it is too complex, the available data and knowledge may be insufficient to accurately parameterize the model and its usefulness will be decreased. Also, models that are too complex can become difficult to interpret. Models may be used in Monte Carlo simulations where measured or estimated distributions of input values are used to generate distributions of output values. This type of modeling has been applied to exposures in aquatic systems (Klaine et al., 1996; ECOFRAM, 1999a; EUFRAM, 2006a) as well as exposures of terrestrial organisms (ECOFRAM, 1999; U.S. EPA, 2001, 2004; EUFRAM, 2006). Output from Monte Carlo simulations is useful for distributional analyses and PRA but, if the model is in error, the error is propagated through the entire data set. Use of Monte Carlo analysis also requires additional information on the distributions of input values, data that may not be available, thus forcing the use of default or assumed values that become a source of systematic errors.

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56.2.2.3  Models for Estimating Exposure   to Pesticides There are a number of models to predict concentrations of pesticides in surface waters and soils. They simulate water flow, sediment deposition, and pollutant fate and transport in surface waters that receive dissolved runoff, soil erosion, and aerial deposition loadings. Outputs include dissolved concentrations of pesticide in the water column and concentrations adsorbed to suspended and bottom sediment as a function of time and location. A widely used model for surface water is EXAMS II (U.S. EPA, 2005a), while PRZM3 (U.S. EPA, 2005b) is used for soils; the two are linked by a PRZM-EXAMS shell (U.S. EPA, 2003). Other models, such as TOXWA (Adriaanse, 1996), which is used in The Netherlands, are available. Models are also available for the FOCUS scenarios used for pesticide risk assessment in Europe (FOCUS, 2001). These models provide estimates of surface and groundwater exposures as well as addressing techniques for mitigation (FOCUS, 2006, 2007a,b). Drift of pesticide sprays is influenced by a number of processes but exposure concentrations through drift are normally reduced with distance from the site of application. There are several approaches to assessing spray drift. One method is to assume that drift is equal to 5% of the application rate (a default assumption previously used by US EPA). Another is to use models such as AGDRIFT, a spray drift model used for aerial applications in the United States (AgDRIFT, 2008). Alternatively, actual measurements of drift resulting from the use of pesticides in the field can be used, as is currently the case in Germany and other parts of Europe (Rautmann et al., 2001). These BBA Drift Tables represent the upper 95th centile drift deposition values from replicate field studies in Europe and are thus reasonable worst-case deposition rates. Although deposition rates may differ in other crops and different locations as well as with the type of application equipment, these can be used as a guideline in areas with similar crops and where similar environmental factors will affect drift. The BBA drift data are incorporated into the FOCUS models discussed above. There are several simplistic models that are used in the initial stages (Tier-1) of risk assessment of pesticides to provide a single value for exposure. The U.S. EPA assumes a water depth of 2 m (based on design specifications for farm ponds) and a direct overspray with the pesticide (SETAC, 1994). In Europe, it is assumed that the water depth of a pond is 100 cm while that in a ditch is 30 cm (FOCUS, 2001). In Canada and the United States, forest pools and wetlands are assumed to be 15 cm deep. Assuming complete and rapid mixing, simple volumetric calculations give worst-case concentrations of 50, 100, 333, and 670  g/l, respectively resulting from a direct surface application of 1 kg/ha. Concentrations from other

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application rates can be determined by simple proportion. Assumptions for the calculation of soil concentrations of pesticides are based on a direct application to soil and mixing in the upper 2.5 or 5 cm of soil. If the pesticide is incorporated (ploughed into) in the soil, mixing is assumed to extend to a depth of 20 cm. If a cover crop is present, it may be assumed to intercept 50% of the spray. Based on a soil density of 1.5 kg/ha, initial concentrations would be 2.67, 1.34, and 0.34 mg/kg resulting from a 1-kg/ha application rate. Interception by the crop will halve the concentration in the soil. Assumptions used in assessing exposure of birds and wildlife are based on estimates used by the FAO of contamination of foliage and seeds and insects (FAO, 1989). These are similar to the Kenaga Nomograms used by the U.S. EPA (Urban and Cook, 1986), as recently reassessed (Fletcher et al., 1994). Small birds (up to 100 g) are assumed to eat a maximum of 30% of their body weight in seeds or insects in a day, while larger birds (500 g) are assumed to consume 10% of their body weight in a day. Based on an application rate of 1 kg/ha, foliage, seeds, large insects, and small insects would be expected to contain 200, 10, 10, and 100 mg/kg wet weight of pesticide, respectively. For a small bird (100 g), the expected dose would be 60, 3, 3, or 30 mg/kg body weight if they ate foliage, seeds, large insects or small insects, respectively. For a large bird (500 g), the dose would be 20, 1, 1, or 10 mg/kg, respectively. There are a number of more refined multicompartment models that can be used to estimate single or multiple values of the pesticide concentration in the environment for use in higher tiers of risk assessment for pesticides. The most simplified of these is GENEEC. GENEEC version 2 (U.S. EPA, 2001b) mimics a PRZM/EXAMS simulation of a generic application of pesticide to field crop for several use patterns and receiving environments. GENEEC is designed as a Tier 1 model. It is conservative and only gives one output value for each use scenario. A more complex combination of EXAMS and PRZM has been used with a preprocessor called the Multiple Scenario Risk Assessment Tool (MUSCRAT, Mangels, 2001). MUSCRAT is an application program that links chemical, crop, soil, and climate data bases (daily values over up to 48 years) and facilitates the creation of PRZM and EXAMS input files, batch processes multiple model simulations, and performs statistical analyses on predicted exposure concentrations for pesticides (ECOFRAM, 1999a). MUSCRAT gives multiple daily values as output and these output values can be analyzed as distributions rather than as single deterministic values. MUSCRAT is designed to provide modeled data for use in higher tiers of PRA. Several other multimedia and probabilistic models for estimating exposures to pesticides are also used. The FOCUS models (FOCUS, 2006, 2007a,b) produce estimates of exposure values but over a shorter time interval (daily over one year). Some Monte Carlo-based models for estimating exposures are integrated

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into risk assessment tools for terrestrial organisms, such as TIM1 and TIM 2 (U.S. EPA, 2001d, 2004b) and for terrestrial and aquatic organisms, WEBFRAM (EUFRAM, 2006b).

56.3  Risk Assessment of Pesticides ERA for pesticides is generally done by comparing the concentration of the pesticide estimated or found in the environmental matrix to response concentrations reported for that pesticide in laboratory toxicity tests. As has been recommended numerous times, risk assessments should be conducted in a series of steps or tiers (SETAC, 1994; ECOFRAM, 1999a,b; Suter et al., 2007). The use of tiered approaches in risk assessment has several advantages. The initial use of conservative criteria allows substances that truly do not present a risk to be eliminated from the risk assessment process, thus allowing the focus of expertise to be shifted to more problematic substances. As one progresses through the tiers, the estimates of exposure and effects become more realistic as uncertainty is reduced through the acquisition of more data. Tiers are normally designed such that the lower tiers in the risk assessment are more conservative (less likely to pass a hazardous chemical), while the higher tiers are more realistic, with assumptions more closely approaching reality. Because lower tiers are designed to be protective, failing to meet the criteria for these tiers is not necessarily an indication of a problem but that more data or a more realistic risk assessment may be needed. Risk assessment of pesticides can be conducted for many reasons. These range from simple ranking systems to more complex PRAs (Figure 56.7).

56.3.1  Scoring Systems and Setting   of Criteria The least complex form of risk assessment is the use of scoring systems to rank substances on the basis of toxicological or other properties. A number of scoring systems

Hayes’ Handbook of Pesticide Toxicology

are used by international and national organizations (IJC, 1993; United Nations, 2007). For example, the World Health Organization (WHO) uses a ranking scheme based on pesticide toxicity to classify pesticides into categories that can be used to regulate the products as well as transportation and storage (World Health Organization, 2004). Many jurisdictions have followed the WHO ranking scheme. The basic principle of a scoring system is to assign a rank or priority to a list of substances. This is usually accomplished by assigning a score to several of the properties of the substances being assessed, processing these scores in some way, and then using the scores to rank (and select) some of these substances for further action or to place these in categories for special handling, storage, etc. Very few scoring systems use decision criteria for multiple values i.e., where different data sources report different values. Most systems use the most conservative (worst-case) value, regardless of source, provenance, or general applicability. Correctly used, scoring systems have been employed to rank substances in order of priority for further assessment. This is usually in the first tier of the risk assessment process where comparisons are needed to assign priorities. Before a final risk management decision, a more detailed risk assessment would be required because the scoring systems rarely consider exposures, commonly make use of worst-case data, and do not handle missing values, weighting, or scaling in clear or appropriate ways. The rank numbers produced from combinations of scores have no meaning in the real world; their only use is to allow prioritization of substances for more detailed assessment. Incorrectly used, scoring systems have been, and are, employed in place of a full risk characterization. This may be politically expedient as it allows rapid action; however, it may result in very controversial or costly decisions being made on the basis of poor science. Toxicity data may be used for setting of environmental criteria or guidelines (Figure 56.1). Criteria and/or guidelines can be derived from worst-case toxicity data (the most sensitive species) with the additional application of an

Figure 56.7  Illustration of the types of hazard and risk assessment from simple scoring systems to probabilistic risk assessment (redrawn from Solomon and Takacs, 2002).

Chapter | 56  Ecotoxicological Risk Assessment of Pesticides in the Environment

uncertainty factor or from distributions of toxicity data to estimate lower centiles of toxicity (see discussion of PRA below). Criteria are usually widely applicable across a jurisdiction, although there are some situations where sitespecific criteria may be developed as well (CCME, 2007). The output of criteria setting is a single number or small range of numbers that are then used to judge if a measured or estimated concentration in the environment is acceptable (will most likely not cause harm) or not (may cause harm). As above for scoring systems, exceeding a criterion does not necessarily mean that harm has occurred. It does, however, require that additional information be sought to determine if an adverse effect has, in fact, occurred.

56.3.2  The Hazard Quotient The first real tier in the risk characterization process is the use of hazard quotients (Figure 56.7). These are simple ratios or quotients of single-exposure and effects values and may be used to express hazard or relative safety. For example: Hazard  Exposure concentration / Effect concentration Margin of Safety  Effect concentration / Exposure concentration



The calculation of hazard quotients is deterministic because it makes use of single values for effect data and exposure data. Normally, the toxicity value for the most sensitive organism or group of organisms is compared to the greatest exposure concentration measured or estimated in the environmental matrix. Note that body-doses of pesticides may be used as well as exposure concentrations, as long as the units of the exposure and the measure of effect are the same. This may be made more conservative by division of the quotient by an uncertainty factor (UF), which may range from 1 to 10,000 (See Section 56.1.3.3, above). This is to allow for unquantified uncertainty in the effect and exposure estimations or measurements. If the hazard quotient (divided by the UF) is greater than 1, a hazard is deemed to exist. In the EU, the margin of safety concept is used instead of the hazard quotient [it is called the toxicity exposure ratio (TER)]. All hazard quotients incorporate some form of uncertainly factor, either explicitly as part of the calculation itself or in the criteria for acceptance of the hazard ratio. In the absence of an adequate range of toxicity tests or exposure measurements, decisions based on hazard quotients may be underprotective and the use of uncertainty factors is justified. However, where an acceptable range of toxicity data is available, the variance in response to the pesticide is better defined and use of safety factors may be overprotective. The quotient approach is thus acceptable for early tiers or preliminary risk assessments but it fails to consider the range of values that may

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exist in terms of exposures and susceptibility when more data are available. As in the case of criteria (above), if the hazard quotient exceeds 1, it may only mean that more data are needed for a more realistic assessment.

56.3.3  Probabilistic Risk Assessment The probability of occurrence of a particular event is, and has been, widely used in the characterization of risk from many physical, medical, and other events in human society (the insurance industry) and for protection against failure in mechanical and civil engineering projects (time between failures, one-in-100-year floods, etc.). This concept has been applied in ecotoxicological risk assessment of pesticides for the characterization of distributions of both exposures and effects. Concentrations of substances in the environment can be affected by a large number of processes that relate to the amount released, the spatial and temporal distributions of the releases, and the results of the action of a large number of transportation and transformation processes (fate processes) on the substance (Table 56.3). The likelihood and extent to which these myriad of fate processes will affect a particular quantity of substance in the environment is essentially random and frequency distributions of exposure concentrations in the environment can be used to describe and characterize the data set and to use it to make predictions similar to those made in other areas of risk management (McBean and Rovers, 1992; Carrington, 1996). In many cases, these distributions best fit the log-normal model (Giddings et al., 2005, to cite but a few references; SETAC, 1994; Klaine et al., 1996; Solomon et al., 1996; Solomon et al., 2001). The assumption of a reasonable fit to a model makes calculations of probabilities that a particular toxicity value will be exceeded relatively easy but is not necessary for the concept of PRA to be used. Centiles of distributions may be estimated from large data sets by simple ranking and interpolation. This is best used with large data sets where extrapolation beyond the observations is not needed. The same observations of log-normality generally apply to distributions of toxicological data. Many of the reactions through which toxicity mechanisms are mediated are first-order or pseudo first-order. Thus, with a large enough data set and appropriate groupings of organisms to avoid mixing susceptible and nonsusceptible species together, good fits to the normal distribution are obtained (Solomon et al., 1996; Solomon and Chappel, 1998; Giesy et al., 1999; Giddings et al., 2000; Kwok et al., 2007). Some care should be taken when using exposure or toxicity data in the distributional analyses. Exposure data should be screened to make sure that the data are consistent. Ideally, exposure data should be expressed over constant intervals, such as in daily samples. Distributional analysis of a data set with unequal time intervals will distort the distribution to over-represent periods where more

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samples were taken. In this case, samples taken more frequently can be combined as time-weighted averages. In situations where samples are taken less frequently, interpolation can be used to “generate” surrogate data with the proviso that this will introduce a bias into the data set. In temperate regions, sampling of environmental matrices may not be carried out in winter. As winter is a period of low biological activity, it may be more ecologically appropriate to focus the risk assessment on the more biologically productive months when more analytical data are available. Effects data used in distributional analyses should be reviewed for appropriate quality (see Section 56.2.1.1, above). However, particularly with older pesticides, multiple studies on the same species that satisfy the appropriate quality criteria may be available. If this is the case, several procedures may be adopted. If one of the data points represents a more sensitive life-stage and life-table analyses indicate that survival of this stage is important for population sustainability, then this datum should be used. If no life-stage data are available and/or multiple tests still satisfy the review criteria, it is recommended that test data be combined as a geometric mean to produce a measure of central tendency in log-normally distributed data (ECOFRAM, 1999a,b). Pesticide toxicity data reported at concentrations in excess of the maximum water solubility of the pesticide may not be reliable descriptors of responses; however, they can be used for risk assessment. These data are almost always from the least susceptible organisms and, while less relevant in the risk assessment, can be used in the characterization of the toxicity distribution. Because the data are less reliable, they should not be used in fitting a model to the distribution but should be included in the calculation of the total sample size (n) and the ranks. The toxicity and exposure data are thus analyzed as distributions on the assumption that the data represent the universe of observations. Obviously, it is not possible to test all the species in an ecosystem and, for this reason, an approximation is usually made and the test data are used as a surrogate. The same is true of exposure data because it is not practical or feasible to sample all possible locations at all possible times. As it is unusual for sufficient toxicity data to be available to allow a cumulative frequency distribution of data to be plotted directly, an approximation is normally used (Parkhurst et al., 1996). The data are ranked and then displayed graphically as log of concentration vs. percent probability. The y-plotting positions (P) are calculated as percentages using the formula:

( P  100  i/ (n  1)) (from Parkhurst et al., 1996),

where i is the rank number of the datum point and n is the total number of data points in the set. This gives and empirical cumulative probability based on the Weibull equation.

Similar empirical probabilities can also be calculated using other formulae such as the Blom equation

( P  (i  0.375) / (n  0.25)  100)

or the Hazen equation

( P  (i  0.5) /n  100),

which are useful for small data sets (10) (Cunnane, 1978). These formulae all compensate for the smaller size of the data set. Small (more uncertain) data sets are more likely to give more conservative estimates of high or low centiles than larger (more certain) data sets. The principle of PRA has been described (Cardwell et al., 1993; SETAC, 1994; Klaine et al., 1996; Parkhurst et al., 1996; Solomon, 1996; ECOFRAM, 1999a,b; Giesy et al., 1999; EUFRAM, 2006a) and is illustrated diagrammatically in Figure 56.8. Distributional analysis can be applied to concentrations of substances in the environment with due consideration for the fact that these data are usually censored by the limits of analytical detection (Figure 56.8A). In practice, all exposure concentration data below the LOD or LOQ are assigned a dummy value of zero. These data are used in the calculation of n for ranking but are not plotted or used to estimate centiles. As in this illustration, when plotted as a cumulative frequency distribution using a probability scale on the y-axis and concentration log10 on the x-axis (Figure 56.8A), this distribution approximates a straight line that can be used to estimate the likelihood that a particular concentration of the substance will be observed, given that the data is a true sample of all the data. A similar approach is used where LC50s or some other toxicity values are plotted (Figure 56.8B) to produce a species sensitivity distribution (SSD; Posthuma et al., 2002). The combination of these distributions in PRA is illustrated in Figure 56.8C. In this procedure, it is assumed that the distributions of sensitivity represent the range of responses that are likely to be encountered in the ecosystems where the exposures occur. If the exposure data were collected over time at a particular site, the degree of overlap of the exposure distribution with the effects distribution can be used to estimate the joint probability of exposure and toxicity. This provides an estimate of the probability that a particular toxicity value or intercept of the SSD curve (i.e., the 10th centile) will be exceeded (Figure 56.8C). If the exposures were measured over time, say days, this is then expressed as proportions of days on which the criterion will be exceeded. This can be applied to a number of data sets and the resulting probabilities used for priority setting or in further assessing ecological relevance. In this example of the process, the probabilities are calculated from the point estimate of the regression line (the cumulative frequency distribution), but they can also be calculated from the lower confidence interval of the toxicity distribution and the upper confidence interval of the exposure

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ecological risk assessments of pollutants (U.S. EPA, 2001c) and pesticides (SETAC, 1994; ECOFRAM, 1999a,b; EUFRAM, 2006a). This approach has been used in a number of risk assessments of pesticides published in the literature (Klaine et al., 1996; Solomon et al., 1996; Solomon and Chappel, 1998; Cardwell et al., 1999; Giesy et al., 1999; Hall et al., 1999; Versteeg et al., 1999; Giddings et al., 2000, 2001; Solomon et al., 2001; Giddings et al., 2005). The major advantage of this approach is that it uses all relevant single species toxicity data and, when combined with exposure distributions, allows quantitative estimations of risks. In addition, the data may be revisited again, the decision criteria become more robust with additional data, and the method is transparent (will give the same results with same data sets). The method does have some disadvantages. More data are usually needed, although these are mostly low-cost studies if acute data are used. For new products, models have to be used to estimate exposures and models may not have been validated for these uses. It is not easily applied to highly bioaccumulative substances where exposure is via food chain as well as matrix; however, if appropriate data, such as body residue values, are available, this can be overcome. Probably most critical of all is that it requires education of risk assessors and risk managers to increase their ability to evaluate decisions and to increase their comfort levels with the process (Solomon, 1996; Bier, 1999; EUFRAM, 2006b; Frewer et al., 2008). PRA can be applied to assessments based on acute or chronic responses; all that is necessary is to ensure that the toxicity data and the exposure data are expressed in the same units. Although more widely used to assess risk to aquatic systems, the techniques are applicable to terrestrial systems as well (ECOFRAM, 1999b; U.S. EPA, 2001d, 2004b; EUFRAM, 2006a).

56.3.3.1  Refinements to the Probabilistic   Risk Assessment Process Figure 56.8  Illustration of the principle of the probabilistic approach (adapted from Solomon and Chappel, 1998). See text for explanation.

distribution. This allows both the variance (range of toxicity or exposure values) and uncertainty about the toxicity and exposure values to be taken into consideration in the risk assessment. Expressing the results of a refined risk assessment as a distribution of values rather than a single point estimate is an approach that has been used in the calculation of water-quality criteria in the United States (Stephan et al., 1985), by the Dutch government (Health Council of the Netherlands, 1993), Australia and New Zealand (ANZECC, 2000), Canada (CCME, 2007), and recommended for use in

A number of refinements to PRA have been suggested by the ECOFRAM working group (ECOFRAM, 1999a) and EUFRAM Process (EUFRAM, 2006a). Grouping organisms on the basis of sensitivity or lack of sensitivity to pesticides, reproductive strategy, and recovery potential is discussed in Section 56.2.1.1, above and is also applicable to PRA. (a) Using the Joint Probability Curve to Characterize Risks Instead of estimating the likelihood that specific toxicity criterion (say the 10th centile of the SSD) will be exceeded, exceedence probabilities can be presented as a continuum of likelihoods in a joint probability curve (JPC) (ECOFRAM, 1999a; EUFRAM, 2006a). The derivation of the JPC (Figure 56.9A, B) is relatively simple and offers

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a useful tool for communication of risks because it allows what-if questions to be addressed and gives the risk assessor and risk manager a method for assessing the effects of changes in assumptions, such as the choice of a different centile from the SSD. The JPC approach has been used to characterize risks in ecological risk assessments of pesticides (Giddings et al., 2005; Giesy et al., 1999; Solomon and Thompson, 2003)

but actual assessment criteria for classifying JPCs have not yet been adopted by the regulatory community. There are some criteria for acceptance of tolerable risks that are, however, used in decision making. Based on an analysis of EPA regulatory practice, Suter et al. (2000) concluded that decreases in an ecological assessment endpoint of less than 20% are generally acceptable. Similar conclusions were stated by Christman et al. (1994) with respect to pond microcosms. In another example, community metrics for an exposed benthic invertebrate community must be reduced by more than 20% compared to pristine reference sites to be considered even slightly impaired in the EPA rapid bioassessment procedure (Barbour et al., 1999). Based on ERAs performed for U.S. EPA at the Calcasieu Estuary, Louisiana (U.S. EPA, 2002), the Housatonic River, Massachusetts (U.S. EPA, 2004a), Moore (D. Moore, 2008, personal communication) has used JPCs to categorize risk to aquatic biota in each acute risk scenario as de minimus, low, intermediate, or high. These are defined below and illustrated as JPCs in Figure 56.9C. If the maximum risk product (risk product  probability of exceedence  magnitude of effect) is 0.25%, then the risk is categorized as de minimus. n If the maximum risk product is 0.25% but 2%, then the risk is categorized as low. n If the maximum risk product is 2% but 10%, then the risk is categorized as intermediate. n If the maximum risk product is 10%, then the risk is categorized as high. n

The curve corresponding to a risk product of 0.25% passes through points corresponding to an extremely small probability of adverse effects (5% probability of 5% affected  0.25%). This value is used as the upper limit for the category of de minimus risk and is consistent with the protection goals of the Dutch HC5 and the U.S. EPA water quality criteria (protection of 95% or more of aquatic species) (Stephan et al., 1985; USEPA, 1995a; Janssen et al., 2004). The curve corresponding to a 2% risk product passes through the points associated with a small probability (10% probability of 20% affected), while the curve corresponding to a risk product of 10% passes through the points associated with a large probability (100% of 10% affected) and a median probability (50% of 20% affected). In the field, scenarios with a maximum risk product of 10% or more are considered to have a large probability of being detectable and possibly cause major impacts on ecosystems. Scenarios with a maximum risk product 2% but 10% are judged to be intermediate risk scenarios.

Figure 56.9  Presentation of exceedence probabilities (A) as a continuum of likelihoods in an exceedence profile/joint probability curve (B) and the suggested use of these curves in decision making (C, adapted from D. Moore, 2008 personal communication).

(b) Probabilistic Risk Assessment for Mixtures Mixtures of pesticides do occur in the environment. Pesticides are rarely used in consistent mixtures or applied in such a way that they will enter and move in the environment to produce predictable combinations of concentrations

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away from the site of application. Because pesticides are regulated as single substances, risk assessments of pesticides have traditionally been conducted on one active ingredient only. The exception to this is mixtures of pesticides that are packaged together or mixed in the spray tank and, here, only acute toxicity testing of the mixture is required. These same constraints apply to PRA. However, environmental monitoring has shown that pesticides do co-occur in the same location and time (USGS, 2007) and this has raised questions regarding risk assessment of mixtures. Where substances are known to act additively, it is possible to use the toxic equivalent (TE) or toxic unit (TU) approach to add concentrations and assess risks from the mixtures. This has been applied to several classes of compounds such as the dioxins (Ahlborg et al., 1994; Parrott et al., 1995; Birnbaum, 1999), chlorinated phenols (Kovacs et al., 1993), and polyaromatic hydrocarbons (Schwarz et al., 1995). This approach has been used to assess the combined risk from atrazine and its metabolites (Solomon, 1999) and triazine herbicides (Giddings et al., 2005) using probabilistic approaches and could be used to assess risks from other pesticides with a common mode of action that are known to act additively on the basis of potency, such as the organophosphorus insecticides. The approach is similar to that used for dioxins in that the concentrations of the components of the mixture are converted to concentration equivalents of one reference component of the mixture, usually the most toxic. These concentrations are then added and compared to toxicity values for the reference compound. Traditionally, these equivalents have been based on responses measured in the same organism, for example the laboratory rat. This is appropriate if the risks are to be assessed in the same organism or extrapolated to another (humans) with appropriate uncertainty factors. However, ratios of potencies measured in one animal may not be the same in another and wide interspecific extrapolations, such as from rats to fish, may not be possible (Parrott et al., 1995). This situation becomes more complex when dealing with ecological risk assessments. If the toxic equivalents are based on responses measured in a single species, the relationship between the potency of the components in the mixture may be different from those in other organisms. Thus, as was the case with dioxins in fish and rats, extrapolation of the risk, whether characterized via the hazard quotient (HQ) or PRA, may be incorrect. Alternatively, TEs could be based on point estimates of potency derived from distributions of toxicity values. Again, use of these TEs would require that the distributions were similar (same slope when expressed as cumulative frequencies curves) and that the order of the species in the distributions is the same. If this is not the case, risk assessments may be incorrect. If the substances interact through response addition, similar approaches would be possible, although differences in exposure times may introduce additional complexity to the assessment.

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If based on measured concentrations from the environmental matrix, PRA of mixtures does have the advantage of being able to provide an assessment of the likelihood of co-occurrence of the substances. This is a considerable improvement over assuming that all the worst-case concentrations will be present in the same location and at the same time. In this case, even though it may be difficult to quantify the total risk in a meaningful way, it would be possible to assess the effect of removal of one of the components of the mixture as a differential risk. This could be useful in risk management decision making. (c) Using Distributional Approaches to Determine Thresholds of Toxicity The no observed effect concentration is widely used in many risk assessments, and where sufficient data are available, may be used in PRA. One concern is that the NOEC is not representative of a concentration at which no biologically significant effect is occurring. A method to address this question and to estimate the threshold of toxicity, or a true NOEC, for aquatic plants has been suggested (Hanson and Solomon, 2002; Brain et al., 2006). The method involves determining the effective concentration (ECx) of a number of endpoints from one species. These ECx values are plotted on a log-probability scale. The x-intercept, or a low centile (such as 0.1%), of the distribution can be interpreted as the threshold of toxicity for that plant (or other organism) at that response level. This threshold is the concentration at which no effects should be observed for any endpoint above that response level and is called the Toxic Benchmark Concentration (TBC; Figure 56.10). It is based on the assumptions that multiple measures of effect from a single species will be log-normally distributed and that the distribution contains all possible endpoints for that species.

Figure 56.10  Distributions of EC values for monochloroacetic acid in Myriophyllum spicatum generated from field-level testing. The horizontal dashed line represents the probability associated with the TBC and the vertical dashed lines represent the threshold of toxicity for the distributions of EC values (adapted from Hanson and Solomon, 2002).

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The thresholds and the distributions can then be used as a substitute for the NOEC or ECx in risk assessment techniques, such as hazard quotients and PERA. The concept has been applied to halogenated acids (Hanson and Solomon, 2002) and pharmaceuticals (Brain et al., 2006).

56.3.3.3  Using Multiple Lines of Evidence in Ecotoxicological Risk Assessment of Pesticides The results of ecological risk assessments need to be interpreted in the context of a number of lines of evidence which include ecological, abiotic, and biotic components of the ecosystem. As currently used, PRA is a purely numerical methodology. This is an advantage from the point of view of the transparency of the procedure, but it cannot, nor is it designed to, assess the ecological relevance of the exceedences of toxicity values that may be identified. For example, an assessment criterion of the 10th centile may include keystone organisms of value to ecosystem function. Effects on keystone species would be expected to extend to other species that are dependent on them, for example as a source of food or as a predator. For this reason, it is necessary to assess the role of the potentially affected species in terms of their function in the ecosystem and whether this can be taken over by other organisms (Hall and Giddings, 2000). The probabilistic approach can be used to refine the assessment process by allowing a rational ranking of scenarios by risk and by identifying species in the distributions for which functional redundancy may exist – less sensitive organisms that can also perform the same function as the more sensitive and more affected organisms. Ecological relevance can most usefully be assessed from a basic knowledge of ecology, which includes functions and roles of species in the ecosystem, from analysis by trait, or trophospecies (Yodzis and Winemiller, 1999; Baird and van den Brink, 2007) and from tests, such as cosms, where community resiliency, productivity, and function can be evaluated directly (van den Brink et al., 2005; van den Brink, 2008). For this reason, refinement of the effects characterization through the use of lines of evidence in a probabilistic risk assessment provides a greater reduction of biological uncertainty. The temporal and spatial scale of pesticide exposure is important in ecological risk assessment. The return frequency of an event (how often the event happens) is an important consideration in the choice of methods for probabilistic risk assessment and is related to the ecological cost of recovery from the event (Solomon, 1996). In assessing exposure, the return frequency protected against should be consistent with the resiliency of vulnerable populations. Resiliency is determined by life cycle characteristics and reproductive capacity of the potentially affected organisms and the ability of their populations (or their function in the ecosystem) to recover from the episode. In temperate regions, many ecosystems undergo a period of

Hayes’ Handbook of Pesticide Toxicology

dormancy and the system is, in a sense, reset seasonally by the winter season. Thus, for some organisms, mechanisms for propagation beyond the winter reset already exist and resting and other dormant stages are produced from which populations in the next season will develop. Similar mechanisms exist in environments with a dry season where ephemeral water bodies are subjected to drying out. Therefore, as many organisms in these regions undergo seasonal resets, a stressor return that occurs less frequently than once per season is likely to be tolerable from the viewpoint of the long-term productivity of the population and the sustainability of function in the ecosystem, especially if the effects are spatially restricted. Protection of longer-lived species without seasonal resets, such as some fish, birds, or mammals, may, however, require the consideration of return frequencies of several years or more. If a stressor is present nonuniformly in the environment, unexposed areas will act as refugia (metapopulations) for repopulation of potentially impacted areas. The relative size of the exposed and unexposed areas and their closeness is important, but this issue is particularly significant for assessing risks from pesticide use, where untreated fields, set-aside land, conservation headlands, crop rotations, and mixed farming practices guarantee that refugia will be present. Similarly, refugia exist in streams and rivers and many organisms have resistant stages or propagules from which population recovery can occur. Thus, probabilistic risk assessments (and hazard quotients) are additionally conservative because they do not consider repopulation from unexposed refugia. The example of the more rapid than expected recovery of the biota in the River Rhine from an endosulfan spill illustrates this point (Friege, 1986). However, recovery in isolated locations, such as ponds, may be very slow in organisms without a protected stage (Blockwell et al., 1999).

56.4  Risk communication As has been pointed out, one of the major hurdles that probabilistic risk assessment will face is its acceptance by the public and regulators (Solomon, 1996; Roberts, 1999; EUFRAM, 2006b; Frewer et al., 2008). Risk managers will likely continue to demand, or at least interpret, probabilistic risk estimates as point estimates with great certainty (Chapman, 1995; Moore and Elliott, 1996; Richardson, 1996). Decision makers want to know whether or not it is safe and prefer being told what will happen, not what might happen (Morgan, 1998). Similarly, the public demands absolute safety but has less understanding of the science and, judging from their addiction to lotteries and other games of chance, has no concept of probability and also greatly misperceives risks to themselves and fellow humans (Slovic et al., 2004, 2005), never mind those to the environment.

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The major reason why acceptance of PRA may be slow is that it reveals and, in some cases, quantifies the uncertainty in risk assessments. This uncertainty is present in all forms of risk assessment but has been obscured by uncertainty factors such as those traditionally applied to hazard quotients. Thus, a strategy is needed for communication between the risk assessor and the risk manager or between the risk manager and politicians and the public. Suggestions for communicating results of PRA have been made (Frewer et al., 2008), but this is not an easy task, especially if the result of the assessment is contrary to conventional wisdom or to the interests of certain stakeholder groups. Written risk communications must be designed to be understood by the lay public and this is a skill that few have. Verbal communication is even more difficult, especially if it is carried out in front of an audience, where nonverbal communication can give mixed signals to the audience. Part of the difficulty with communicating risk assessment is that, by human definition, risk is assumed to be adverse. In the final analysis, in the ecosystem there is neither “good” nor “bad” and certainly nothing “adverse”. An ecological change is labeled “adverse” by individuals or society and is basically a value judgment. Thus, to conduct a risk assessment means that someone has made a value judgment of which conditions will be defined as adverse. The public is often suspicious of the motives of those who communicate risks and may perceive conflicts of interest (Lackey, 1995). Risk communication is a form of persuasive communication that is designed to change behavior or what may be deeply held beliefs. Two of the keys to successful risk communication are (1) expressing the technical evaluation in a way that is meaningful to the audience, which can be achieved by using appropriate analogies to describe the risk assessment process and not using technical terminology that can be misunderstood and (2) anticipating potential misunderstandings and dealing with them in a sympathetic way. In doing this, the communicator must recognize that groups that oppose a particular risk management strategy have a right to question a decision that affects them.

Conclusion The environmental risk assessment of pesticides has seen significant advances in the last decade. Many of these advances have been related to the development of probabilistic approaches for the characterization of data for exposures and toxicity, thus better defining variability in these key components of risk assessment. Methods for characterizing uncertainty are being developed and will result in further evolution of the assessment process. However, risk assessment procedures, such as PRA, and the more traditional HQ are numerical and do not implicitly consider

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ecological interactions and the role of species in the ecosystem. For example, occasional exceedences of thresholds of toxicity may have no long-term direct effects on a species with a high potential for recovery but the same may not be true of a species with a long generation time or low potential for recovery. Similarly, keystone species and their role in ecosystems must be considered. Thus, basic biological and ecological knowledge must be retained in the risk assessment process if the significance of ecological risks is to be properly considered.

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Virtual Elimination Task Force to the IJC. International Joint Commis­ sion, Windsor, ON, Canada. James, F. C. (1994). Society actions. Joint EAS/AIBS review of President Clinton’s plan for the management of forests in the Pacific Northwest. Bull. Ecol. Soc. Am. June, 69–75. Janssen, M. P. M., Traas, T. P., Rila, J.-P., and van Vlaardingen, P. L. A. (2004). Guidance for Deriving Dutch Environmental Risk Limits from EU-Risk Assessment Reports of Existing Substances. Bilthoven, NL, RIVM, Rep. No. 601501020/2004, pp 35, http://www.rivm.nl/ bibliotheek/rapporten/601501020.pdf. Klaine, S. J., Cobb, G. P., Dickerson, R. L., Dixon, K. R., Kendall, R. J., Smith, E. E., and Solomon, K. R. (1996). An ecological risk assessment for the use of the biocide, dibromonitrilopropionamide (DBNPA) in industrial cooling systems. Environ. Toxicol. Chem. 15, 21–30. Kovacs, T. G., Martel, P. H., Voss, R. H., Wrist, P. E., and Willes, R. F. (1993). Aquatic toxicity equivalency factors for chlorinated phenolic compounds present in pulp mill effluent. Environ. Toxicol. Chem. 12, 684–691. Kwok, K. W. H., Leung, K. M. Y., Chu, V. K. H., Lam, P. K. S., Morritt, D., Maltby, L., Brock, T. C. M., van den Brink, P. J., Warne, M. S. J., and Crane, M. (2007). Comparison of tropical and temperate freshwater species sensitivities to chemicals: implications for deriving safe extrapolation factors. Integr. Environ. Assess. 3, 49–67. Lackey, R. T. (1995). The future of ecological risk assessment. Human Ecol. Risk Assess. 1, 339–343. Lehtinen, K.-J., Axelsson, B., Kringstad, K., and Strömberg, L. (1991). Characterization of pulp mill effluents by the model ecosystem technique: SSVL investigations in the period 1982–1990. Nordic Pulp Pap. Res. J. 2, 81–88. Liber, K., Solomon, K. R., Kaushik, N. K., and Carey, J. H. (1994). Impact of 2,3,4,6-tetrachlorophenol (DIATOX) on plankton communities in limnocorrals. In “Aquatic Mesocosm Studies in Ecological Risk Assessment,” (R. L. Graney, J. L. Kennedy, and J. H. Rogers, eds.), pp. 257–294. Lewis Publishers, Boca Raton, FL. Liess, M., Brown, C., Dohmen, P., Duquesne, S., Hart, A., Heimbach, F., Krueger, J., Lagadic, L., Maund, S., Reinert, W., Streloke, M., and Tarazona, J. V.(2005). “Effects of Pesticides in the Field. (EPIF),” pp. 136. SETAC Press, Brussels, Belgium. Maltby, L., Blake, N. N., Brock, T. C. M., and van den Brink, P. J. (2005). Insecticide species sensitivity distributions: The importance of test species selection and relevance to aquatic ecosystems. Environ. Toxicol. Chem. 24, 379–388. Mangels, G. (2001). “The Development of MUSCRAT (Multiple Scenario Risk Assessment Tool): A Software Tool for Conducting Surface Water Exposure Assessments,” Accessed Jan 2, 2007 http://www. waterborne-env.com. Waterborne Environmental, Inc. McBean, E. A., and Rovers, F. A. (1992). Estimation of the probability of exceedences of a contaminant concentration. Ground Wat. Monitor. Rev. 12, 115–119. Mendelson, J. R. III, Lips, K. R., Gagliardo, R. W., Rabb, G. B., Collins, J. P., Diffendorfer, J. E., Daszak, P., Ibáñez D, R., Zippel, K. C., Lawson, D. P., Wright, K. M., Stuart, S. N., Gascon, C., da Silva, H. R., Burrowes, P. A., Joglar, R. L., La Marca, E., Lötters, S., du Preez, L. H., Weldon, C., Hyatt, A., Rodriguez-Mahecha, J. V., Hunt, S., Robertson, H., Lock, B., Raxworthy, C. J., Frost, D. R., Lacy, R. C., Alford, R. A., Campbell, J. A., Parra-Olea, G., Bolaños, F., Domingo, J. J. C., Halliday, T., Murphy, J. B., Wake, M. H., Coloma, L. A., Kuzmin, S. L., Price, M. S., Howell, K. M., Lau, M., Pethiyagoda, R., Boone, M., Lannoo, M. J., Blaustein, A. R., Dobson, A., Griffiths, R. A., Crump, M. L., Wake, D. B., and Brodie, E. D. Jr. (2006). Confronting amphibian declines and extinctions. Science 313, 48.

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Moore, D. R. J., and Elliott, B. J. (1996). Should uncertainty be quantified in human and ecological risk assessments used for decision-making?. Human Ecol. Risk Assess. 2, 11–24. Morgan, M. G. (1998). Uncertainty analysis in risk assessment. Human Ecol. Risk Assess. 4, 25–39. OECD (1998). “The OECD Principles of Good Laboratory Practice,” Rep. No. Environment Monograph No. 45, pp. Organisation for Economic Co-Operation and Development, Paris, France. OECD (2002). “Manual for Investigation of HPV Chemicals, Chapter 4: Initial Assessment of Data,” Rep. No. Guidance currently being finalized, pp. Organisation for Economic Co-Operation and Development, Paris, France. OECD (2008). OECD Guidelines for the Testing of Chemicals. Section 2: Effects on Biotic Systems. OECD, Accessed January 31, 2008 http:// miranda.sourceoecd.org/vl3332085/cl13/nw1/rpsv/cw/vhosts/ oecdjournals/1607310x/v1n2/contp1-1.htm. Parkhurst, B. R., Warren-Hicks, W. J., Cardwell, R. D., Volosin, J. S., Etchison, T., Butcher, J. B., and Covington, S. M. (1996). “Aquatic Ecological Risk Assessment: A Multi-Tiered Approach to Risk Assessment,” Rep. No. 91-AER-1. Water Environment Research Foundation, Alexandria, VA, USA. Parrott, J. L., Hodson, P. V., Servos, M. R., Huestis, S. L., and Dixon, D. G. (1995). Relative potencies of polychlorinated dibenzo-p-dioxins and dibenzofurans for inducing mixed function oxygenase activity in rainbow trout. Environ. Toxicol. Chem. 14, 1041–1050. Posthuma, L., Suter, G. W., and Traas, T. (2002). Species sensitivity distributions in risk assessment. In Species Sensitivity Distributions in Ecotoxicology, p. 564. CRC Press, Boca Raton, FL, USA. Rautmann, D., Streloke, M., and Winkler, R. (2001). New basic drift values in the authorisation procedure for plant protection products. In “Workshop on Risk Assessment and Risk Mitigation Measures in the Context of the Authorization of Plant Protection Products (WORMM)” (R. Forster and M. Streloke, Eds.), Vol. 381. Mitt. Biol. Bundesanst. Land-Forstwirtsch, Berlin, Germany. Reinert, K. H., Bartell, S. M., and Biddinger, G. R. (1998). “Ecological Risk Assessment Decision-Support System: A Conceptual Design,” (C. G. Ingersol, Ed.), pp. 98. SETAC Press, Pensacola. Richardson, G. M. (1996). Deterministic versus probabilistic risk assessment: Strengths and weaknesses in a regulatory context. Human Ecol. Risk Assess. 2, 44–54. Roberts, S. M. (1999). Practical issues in the use of probabilistic risk assessment. Human Ecol. Risk Assess. 5, 729–736. Schwarz, R. C., Schults, D. W., Ozretich, R. W., Lamberson, J. O., Cole, F. A., DeWitt, T. H., Redmond, M. S., and Ferraro, S. P. (1995). Sigma PAH: A model to predict the toxicity of polynuclear aromatic hydrocarbon mixtures in field-collected sediments. Environ. Toxicol. Chem. 14, 1977–1978. SETAC (1994). “Society for Environmental Toxicology and Chemistry. Pesticide Risk and Mitigation. Final Report of the Aquatic Risk Assessment and Mitigation Dialog Group,” SETAC Foundation for Environmental Education, Pensacola, FL, USA. Slovic, P., Finucane, M. L., Peters, E., and MacGregor, D. G. (2004). Risk as analysis and risk as feelings: some thoughts about affect, reason, risk, and rationality. Risk Anal. 24, 311–322. Slovic, P., Peters, E., Finucane, M. L., and Macgregor, D. G. (2005). Affect, risk, and decision making. Health Physiol. 24, S35–S40. Solomon, K. R. (1996). Overview of recent developments in ecotoxicological risk assessment. Risk Anal. 16, 627–633. Solomon, K. R. (1999). Integrating environmental fate and effects information: the keys to ecotoxicological risk assessment for pesticides.

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In “Pesticide Chemistry and Bioscience: The Food-Environment Challenge,” (G. T. Brooks and T. R. Roberts, eds.), pp. 313–326. Royal Society of Chemistry, London, UK. Solomon, K. R. and Chappel, M. J. (1998). Triazine herbicides: ecological risk assessment in surface waters. In Triazine Risk Assessment (L. Ballantine, J. McFarland, and D. Hackett, eds.) Vol. 683, pp. 357–368. American Chemical Society, Washington, DC, USA. Solomon, K. R. and Thompson, D. G. (2003). Ecological risk assessment for aquatic organisms from over-water uses of glyphosate. J. Toxicol. Environ. Hlth. B. 6, 211–246. Solomon, K. R., Baker, D. B., Richards, P., Dixon, K. R., Klaine, S. J., La Point, T. W., Kendall, R. J., Giddings, J. M., Giesy, J. P., Hall, L. W. J., Weisskopf, C., and Williams, M. (1996). Ecological risk assessment of atrazine in North American surface waters. Environ. Toxicol. Chem. 15, 31–76. Solomon, K. R., Brock, T. C. M., de Zwart, D., Dyer, S. D., Posthuma, L., Richards, S. M., Sanderson, H., Sibley, P. K., and van den Brink, P. J. (2008a). Extrapolation for criteria setting and risk assessment. In “Extrapolation Practice for Ecotoxicological Effect Characterization of Chemicals,” (K. R. Solomon, T. C. M. Brock, D. de Zwart, S. D. Dyer, L. Posthuma, S. M. Richards, H. Sanderson, P. K. Sibley, and P. J. van den Brink, eds.), pp. 1–32. CRC/Taylor and Francis/SETAC Press, Pensacola, FL, USA. Solomon, K. R., Brock, T. C. M., De Zwart, D., Dyer, S. D., Posthuma, L., Richards, S. M., Sanderson, H., Sibley, P. K., and van den Brink, P. J. (2008b). “Extrapolation Practice for Ecotoxicological Effect Chara­ cterization of Chemicals,” pp. 368. CRC/Taylor and Francis/SETAC Press, Pensacola, FL, USA. Solomon, K. R., Giddings, J. M., and Maund, S. J. (2001). Probabilistic risk assessment of cotton pyrethroids: I. Distributional analyses of laboratory aquatic toxicity. Environ. Toxicol. Chem. 20, 652–659. Stephan, C. E., Mount, D. I., Hansen, D. J., Gentile, J. H., Chapman, G. A., and Brungs, W. A. (1985). “Guidelines for Deriving Numerical National Water Quality Criteria for the Protection of Aquatic Orga­ nisms and their Uses,” Rep. No. PB 85-227049, pp 97. United States EPA ORD ERL, Duluth, MN, USA. Stephenson, G. L., Kaushik, N. K., Solomon, K. R., Day, K. E., and Hamilton, P. (1986). Impact of methoxychlor on freshwater plankton communities in limnocorrals. Environ. Toxicol. Chem. 5, 587–603. Suter, G. W., Barnthouse, L. W., Bartell, S. M., Mill, T., Mackay, D., and Patterson, S. (1993). “Ecological Risk Assessment,” Lewis Publishers, Boca Raton, FL. Suter, G. W. II, Efroymson, R. A., Sample, B. E., and Jones, D. S. (2000). “Ecological Risk Assessment for Contaminated Sites,” Lewis Publishers, Boca Raton, FL, USA. Suter, G. W. II, Barnthouse, L. W., Bartell, S. M., Cormier, S. M., Mackay, D., Mackay, N., and Norton, S. B. (2007). “Ecological Risk Assessment,” CRC Press/Taylor and Francis, Boca Raton, FL. Tilman, D. (1996). Biodiversity: Population versus ecosystem stability. Ecol. 77, 350–363. Tilman, D., Wedlin, D., and Knops, J. (1996). Productivity and sustainability influenced by biodiversity in grassland ecosystems. Nature 379, 718–720. United Nations (2007). “Globally Harmonized System of Classification and Labelling of Chemicals (GHS),” Rep. No. ST/SG/AC.10/30/ Rev.2, pp 555, http://www.unece.org/trans/danger/publi/ghs/ghs_ rev02/02files_e.html. United Nations, New York and Geneva. Urban, D. J. and Cook, N. J. (1986). “Standard Evaluation Procedure for Ecological Risk Assessment,” Rep. No. EPA/540/09-86/167. Hazard Evaluation Division, Office of Pesticide Programs, United States Environmental Protection Agency, Washington, DC.

Chapter | 56  Ecotoxicological Risk Assessment of Pesticides in the Environment

U.S. EPA (1992). “Framework for ecological risk assessment,” Rep. No. EPA/630/R-92/001, pp 41. United States Environmental Protection Agency, Washington, DC, USA. U.S. EPA (1995a). Final Water Quality Guidance for Great Lakes System. In “Fed. Reg.” pp. 15366–15425, Washington, DC, USA. U.S. EPA (1995b). “The use of the benchmark dose approach in health risk assessment,” Rep. No. EPA/630/R-94/007, pp 77. United States Environmental Protection Agency, Washington, DC, USA. U.S. EPA (1998). “Guidelines for Ecological Risk Assessment,” 191. United States Environmental Protection Agency, Washington, DC, USA. U.S. EPA (2001a). Environmental effects database (EEDB). In ECOTOX Database System. United States Environmental Protection Agency, Office of Pesticide Programs, Environmental Fate and Effects Division, United States EPA, Washington, D.C. Accessed June 10, 2006 http://www.epa.gov/ecotox/ U.S. EPA (2001b). “GENEEC – (GEN)eric (E)stimated (E)nvironmental (C)oncentration Model,” http://www.epa.gov/oppefed1/models/water/ geneec2_users_manual.htm. USA, Environmental Fate and Effects Division, Office of Pesticide Programs, USEPA, Washington, DC. U.S. EPA (2001c). “Risk Assessment Guidance for Superfund: Volume III Part A – Process for Conducting Probabilistic Risk Assessment,” Rep. No. 540-R-02-002, pp 385http://www.epa.gov/oswer/riskassessment/ rags3adt/pdf/rags3adt_complete.pdf. USA, Office of Emergency and Remedial Response, United States Environmental Protection Agency, Washington, DC. U.S. EPA (2001d). “Terrestrial Investigation Model (TIM V1),” http:// www.epa.gov/oppefed1/ecorisk/presentations/eufram_overview.htm. USEPA EFED, Washington, DC, USA. U.S. EPA (2002). “Calcasieu Estuary Remedial Investigation/Feasibility Study (RI/FS): Baseline Ecological Risk Assessment (BERA),” Region 6, 37, http://mapping2.orr.noaa.gov/portal/calcasieu/calc_ html/pdfs/reports/berassessws.pdf. United States Environmental Protection Agency, Dallas, TX, USA. U.S. EPA (2003). “PRZM-EXAMS shell PE4V01.pl,”Accessed December 1, 2008 http://www.epa.gov/oppefed1/models/water/pe4memo.htm. United States EPA, Office of Research and Development. U.S. EPA (2004a). “Ecological Risk Assessment for General Electric (GE)/Housatonic River Site, Rest of River,” http://www.epa.gov/ region1/ge/thesite/restofriver/reports/era_nov04/215498_ERA_ FNL_TOC_MasterCD.pdf. United States Environmental Protection Agency, New England Region, Boston, MA, USA. U.S. EPA (2004b). “Terrestrial Investigation Model (TIM V2),” http:// www.epa.gov/oppefed1/ecorisk/presentations/eufram_overview.htm. USEPA EFED, Washington, DC, USA. U.S. EPA (2005a). “Exposure Analysis Modeling System: EXAMS II, Version 2.98.04.06,” Accessed December 1, 2008 http://www.epa. gov/ceampubl/swater/exams/exams2980406.htm. United States EPA, Office of Research and Development. U.S. EPA (2005b). “PRZM3 A Model for Predicting Pesticide Fate in the Crop Root Zone and Unsaturated Soil Zones. Version 3.12.3,” Accessed December 1, 2008 http://www.epa.gov/ceampubl/gwater/przm3/przm3122.htm. United States Environmental Protection Agency. U.S. EPA (2007a). “Endocrine Disruptor Screening Program Chronology,” Accessed December 28, 2007 http://www.epa.gov/scipoly/oscpendo/

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pubs/edspoverview/chronology.htm. United States Environmental Protection Program. U.S. EPA (2007b). “Good Laboratory Practice Standards,” Accessed December 2, 2008 http://www.access.gpo.gov/nara/cfr/waisidx_ 07/40cfr160_07.html. United States Environmental Protection Agency. USGS (2007). “NAWQA Database,” Accessed July 2, 2007 http:// infotrek.er.usgs.gov/servlet/page?_pageid543&_dadportal30&_ schemaPORTAL30. United States Geological Survey. van den Brink, P. J., (2008). Ecological risk assessment: From bookkeeping to chemical stress ecology. Environ. Sci. Technol. 42, 8999–9004. van den Brink, P. J., and Ter Braak, C. J. F. (1999). Principal response curves: analysis of time-dependent multivariate responses of biological community to stress. Environ. Toxicol. Chem. 18, 138–148. van den Brink, P. J., Tarazona, J. V., Solomon, K. R., Knacker, T., van den Brink, N. W., Brock, T. C. M., and Hoogland, J. P. H. (2005). The use of terrestrial and aquatic microcosms and mesocosms for the ecological risk assessment of veterinary medicinal products. Environ. Toxicol. Chem. 24, 820–829. van den Brink, P. J., Sibley, P. K., Ratte, H. T., Baird, D. J., Nabholz, J. V., and Sanderson, H. (2008). Extrapolation of Effects Measures across Levels of Biological Organization in Ecological Risk Assessment. In “Extrapolation Practice for Ecotoxicological Effect Characterization of Chemicals,” (K. R. Solomon, T. C. M. Brock, D. de Zwart, S. D. Dyer, L. Posthuma, S. M. Richards, H. Sanderson, P. K. Sibley, and P. J. van den Brink, eds.), pp. 105–133. CRC/Taylor and Francis/ SETAC, Pensacola, FL, USA. van Straalen, N. M. (2003). Ecotoxicology becomes stress ecology. Environ. Sci. Technol. 37, 324A–330A. Versteeg, D. J., Belanger, S. E., and Carr, G. J. (1999). Understanding single-species and model ecosystem sensitivity: Data-based comparison. Environ. Toxicol. Chem. 18, 1329–1346. Walker, B. (1992). Biodiversity and ecological redundancy. Conserv. Biol. 6, 18–23. Walker, B. (1995). Conserving biological diversity through ecosystem resilience. Conserv. Biol. 9, 747–752. Wheeler, J. R., Leung, K. M. Y., Morritt, D., Sorokin, N., Rogers, H., Toy, R., Holt, R., Whitehouse, P., and Crane, M. (2002). Freshwater to saltwater toxicity extrapolation using species sensitivity distributions. Environ. Toxicol. Chem. 21, 2459–2467. Wilson, C. J., Brain, R. A., Sanderson, H., Johnson, D. J., Bestari, K. T., Sibley, P. K., and Solomon, K. R. (2004). Structural and functional responses of plankton to a mixture of four tetracyclines in aquatic microcosms. Environ. Sci. Technol. 38, 6430–6439. World Health Organization(2004). “The WHO recommended classification of pesticides by hazard and guidelines to classification,” 56, http://www.inchem.org/documents/pds/pdsother/class.pdf. World Health Organization, Geneva, Switzerland. Yodzis, P. and Winemiller, K. O. (1999). In search of operational trophospecies in a tropical aquatic food web. Oikos 87, 327–340. Zeeman, M. and Gilford, J. (1993). Ecological hazard evaluation and risk assessment under EPA’s Toxic Substances Control Act (INV): An introduction. In “Environmental Toxicology and Risk Assessment,” (W. G. Landis, J. S. Hughes, and M. A. Lewis, eds.), pp. 7–21. American Society for Testing and Materials, Philadelphia, PA.

Chapter 57

Environmental Transport and Fate James N. Seiber and Loreen Kleinschmidt University of California, Davis, California

The approaches taken from roughly the 1970s to the present toward understanding the principles of pesticide transport and fate in the environment led to development of a prospective, predictive capability for evaluating environmental behavior before widespread use or release. Progress has been made in defining and understanding dissipation pathways, the relationship between physicochemical properties and dissipation, structure–activity relationships, and environmental activation and deactivation. Improvements in analytical methodology, which provide much lower detection limits for following the fate of breakdown products as well as parent chemicals, and modeling of processes and their effects have been central to the development of a principle-based approach to pesticide processing in the environment.

57.1  Introduction Assessing the transport and fate of pesticides in the envir­ onment is complicated. There are many environmental pathways available at the local, regional, and global levels. Pesticides vary greatly in physical and chemical properties and use patterns, plus the environment itself is complex and varies from one location to another and from one time to another. It is a goal of environmental sciences to understand and deal with the complexities in nature by defining and sorting out underlying principles. These can serve as a basis for developing an assessment of chemical processing and its relationship to the health of the environment. In the past, including the first decades after the widespread introduction of synthetic organic chemicals for pest control, knowledge of environmental behavior and fate was determined by analysis for these chemicals in environmental samples after they had been used/released for many years. By analyzing soil, water, sediment, air, plants, and animals, environmental scientists were able to piece together profiles of each chemical’s environmental Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

behavior. Dibromochloropropane, ethylene dibromide, and chemicals with similar uses as soil nematicides and similar physical/chemical properties were recognized for their potential to contaminate groundwater in general use areas. DDT (dichlorodiphenyltrichloroethane) and other chlorinated insecticides and organic compounds of similar low polarity and water solubility and high stability threatened some aquatic and terrestrial organisms because of their potential for undergoing bioaccumulation and their chronic toxicities. Like the chlorofluorocarbons, methyl bromide was found to be quite stable in the atmosphere and able to diffuse to the stratosphere, where it entered into reaction sequences contributing to the thinning of the ozone layer. However, the retrospective approach is fraught with difficulty: 1. Adverse chemical behavior might be discovered too late, after considerable environmental damage was already done. Examples are the decline of raptorial bird species after widespread use of DDT and substantial loss of stratospheric ozone from long-term use and release of chlorofluorocarbons and methyl bromide. 2. By analyzing for the wrong chemical or the wrong target media, the problem may be misdefined or completely overlooked. For example, parent pesticides such as aldicarb or aldrin appear to have low persistence in the environment, but they can be converted into breakdown products (aldicarb sulfoxide and sulfone; dieldrin and, eventually, photodieldrin), which may be the primary offenders. Targeting only the parents rather than the products in the analysis scheme may overlook the more hazardous products. The trend from roughly the 1970s to the present has been to develop an understanding of the underlying principles of environmental fate to find ways to predict environmental behavior before the chemical is released. For economic materials such as pesticides, premarket environmental fate/ effects testing is now built into the requirements for regulatory approval. The Environmental Fate Guidelines of 1219

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the U.S. Environmental Protection Agency (EPA) (Barrett and Behl, 2003; Kovacs, 1983; U.S. EPA, 1982, 2008) were developed to specify the tests and acceptable behavior required for registration of candidate pesticides in the United States. In order to minimize variations among testing procedures required under different U.S. laws, EPA’s Office of Prevention, Pesticides and Toxic Substances (OPPTS) now recommends that pesticide regis­trants provide data from tests conducted according to the EPA’s OPPTS Harmonized Test Guidelines. Updated guidelines for harmonized environmental fate tests are given at the EPA website (U.S. EPA, 2009b). In the future, harmonization efforts will continue on the interagency level, via the U.S. Interagency Coordinating Committee for the Validation of Alternative Methods (ICCVAM) (U.S. EPA, 2009a). Similar guidelines and test protocols were developed in Europe (Thomas, 1991), Canada (Agriculture Canada, 1987), Australia (Holland, 1999), and other nations and economic organizations. In an effort to reduce the burden on chemical producers, promote efficient use of scientific resources, and form a basis for cooperation, member nations of the Organisation for Economic Co-operation and Development (OECD) have developed internationally harmonized data requirements and standard test methods (OECD, 2009). Another stimulus for better analytical and predictive tests was the development of risk assessment tools for evaluating risks of chemicals in the environment. Risk assessment and risk science in general for the evaluation of health impacts of chemicals date from the late 1970s and early 1980s for human health risk assessment (National Research Council, 1983) and even more recently for ecological risk assessment (Suter et al., 1993). For use in both the hazard identification component, which includes measuring/estimating emissions to the environment, and, particularly, the exposure assessment component of risk assessment, which involves measuring or modeling exposures via food, water, air, etc. (Krieger and Woodrow, 2007), predictive tools (models) are under­going rapid development for use in regulatory actions, both for premarket screening and for reaching decisions on continuing use. Many commercial pesticides, as well as hazardous air pollutants (National Research Council, 1994) and other chemicals of environmental concern, have undergone or are now in the process of undergoing risk assessment review (National Research Council, 1993). Much of the spotlight has been focused on chemicals already in use or present in the environment to guide societal decisions on what steps might need to be taken to reduce adverse impacts. However, the methodology can also be applied to simulate hazards and risks for new candidate pesticides. Both the regulatory agencies and industry have played important roles in encouraging the development of predictive methods. It is clearly in the best interests of companies to screen out potential environmental problems early in the

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development process and to focus resources on chemicals with prospects for long-term environmental compatibility. For example, environmental scientists at Dow Chemical in the early 1970s developed a “benchmark approach” for evaluating the environmental characteristics of candidate pesticides (Goring and Hamaker, 1972). The benchmark approach and other early developments in screening/predicting environmental behavior including modeling became formalized in the field of environmental chemodynamics, which may be generally defined as “the subject dealing with the transport of chemicals (intra-and interphase) in the environment, the relationship of their physical–chemical properties to transport, their persistence in the biosphere, their partitioning in the biota, and toxicological and epi­ demiological forecasting based on physicochemical properties” (Haque and Freed, 1974; Thibodeaux, 1979). Another factor in developing a predictive capability for environmental behavior and fate is the shift in the types of pesticide chemicals over the past 40 years or so. The highly stable, lipophilic organochlorines, organophosphates of relatively high mammalian toxicity, and environmentally persistent triazine and phenoxy herbicides that dominated pesticide chemistry in the 1970s either are gone entirely from pesticide markets or are being replaced. In their place are synthetic pyrethroids, neonicotinoids, sulfonylureas, aminophosphonic acid derivatives, biopesticides, and many other classes and types whose environmental fate and ecotoxicological effects are less straightforward. Some of the new pesticides are attractive because they degrade relatively rapidly and extensively in the environment, but this can multiply the number of discrete chemicals that need to be evaluated. Relying solely on experimentation under environmental conditions could significantly slow regulatory approval, another argument for using predictive screening assessment tools as an integral component of the overall research and development effort. There is increasing pressure to develop tests for subtle environmental effects that go beyond the persistence, leaching, bioaccumulation, and acute/chronic toxicity testing prominent in environmental fate tests in the past. Environmental endocrine disruption effects caused by trace levels of chemicals and chemical mixtures represent an example (Inter-Organization Programme for the Sound Management of Chemicals, 1997; Jobling et al., 1995; National Research Council, 1999; Younglai et al., 2006). Ideally, environmental chemists would be able to detect interactions of endocrine-disrupting chemicals (EDCs) with mammalian tissues and ecosystems by biobased testing for the chemicals themselves or biomarkers that indicate that exposures to EDCs had occurred. The methods of and approaches to screening for EDCs, under intense development from the stimulus of the Food Quality Protection Act (FQPA, 1996; Johnson and Bailey, 1999), add significantly to an already complex array of premarket predictive tests.

Chapter | 57  Environmental Transport and Fate

Much of our current capability for determining transport and fate of pesticides and other chemicals may be traced to the tremendous developments in analytical chemistry from the 1950s and 1960s to the present. Detection limits of low parts per billion (ppb) and even parts per trillion (ppt) are now achievable by better methods of extracting, preparing, and, particularly, determining residues of pesticides and breakdown products in a variety of matrices (e.g., Fong et al., 1999; Seiber, 2005). Developments in gas and liquid chromatography, mass spectrometry, and immunoassay have been among those most useful to envir­ onmental scientists, but computer data handling capabilities have also enabled the routine use of these sophisticated techniques in both research and regulatory laboratories. Biosensor-based methods provide promise for superior detectability in miniaturized equipment that can be used in the field (Van Emon, 2007), and new methods for rapidly extracting and fractionating samples have simplified and increased throughput of laboratory analyses that were often backlogged and slow to deliver useable results (Lehotay, 2006). New developments such as metabolomics enable snapshots of all the important chemicals present in complex mixtures with little or no sample preparation, cleanup, or fractionation. There exists a growing number of applications for these sophisticated analyses in living tissues, biofluids, excreta, and other sample types (Dixon et al., 2006).

57.2  Principles 57.2.1  The Dissipation Process Once a substrate (agricultural commodity, body of water, wildlife, soil, etc.) has been exposed to a chemical intentionally or accidentally, dissipation processes begin immediately. The initial residue dissipates at an overall rate that is a composite of the rates of volatilization, washing off, leaching, hydrolysis, microbial degradation, and other individual processes (Crosby, 1998; Seiber, 1985). Concentrations of the total residue (parent chemical plus breakdown products) typically decrease with time after exposure or treatment ends. For chemicals that are converted to products that are more stable than the parent, product formation can slow the overall dissipation process, well exemplified by significant concentrations of dichlorodiphenyldichloroethane (DDD) and dichlorodiphenyldichloroethylene (DDE) in field soils exposed to DDT many years previously. When low-level exposure results in the accumulation of residues over time, as in the case of bioconcentration of residues from water by aquatic organisms, the overall environmental process includes both the accumulation and dissipation phases. However, for simple dissipation, such as that occurring in the application of pesticides and resulting exposure from residues in food or water or air, the typical results are that concentrations of

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overall residue (parents plus products) decrease with time after exposure to treatment (Figure 57.1). Because most individual dissipation processes follow apparent firstorder kinetics, overall dissipation or decline is also often observed to be first order. Because first-order decline processes are logarithmic, that is, a plot of remaining residue concentration versus time is asymptotic with respect to the time axis, residues will approach zero with time but never cease to exist entirely (Figure 57.1a). Thus all environmental exposures lead to residues that have, theoretically, unlimited residue longevity. But our ability to detect those residues is limited by the operational abilities of analyt­ical methods of gas chromatography, high-performance liquid chromatography, mass spectrometry, immunoassay, and other approaches. The goal is to have sufficient detectability in the methods used to be able to follow residues to the point where their amounts are well below any plausible threshold for adverse biological effects. This presents an inherent dilemma, because biological effects testing is subject to continuing reevaluation (e.g., with environmental endocrine disruptors). Thus, more sensitive analytical techniques are constantly being developed so that dissipation processes can be followed to lower concentration levels, with more chemical breakdown product detail.

57.2.2  Environmental Compartments Once a pesticide gains entry to the environment, by intentional or accidental release, it may enter one or more compartments, illustrated in Figure 57.2. The initial compartment contacted by the bulk of the pesticide will be determined largely by the manner of use or release. Pesticides applied to flooded rice fields, for example, enter the aquatic compartment initially. In time, however, residues will tend to redistribute and favor one or more compartments or media over others, in accord with the chemicals’ physical properties, reactivity, and stability characteristics and the availability and composition of compartments in the general environment where the use or release has occurred (Biggar and Seiber, 1987). Figure 57.2 tabulates the compartments, the transfer/transformation process, and the environmental characteristics that are involved in transport and fate. The nature of the chemical of interest will dictate the pathways to be favored, so that environmental dissipation and fate must be evaluated on a ­ chemical by chemical basis, as well as an environment-­specific basis. This is illustrated in Figure 57.3 for chemical behavior in a pond environment, for which the properties of the chemical of interest must be taken into account along with, and as influenced by, the properties of the pond environment. Analogous schematics have been developed for chemical behavior in soil (Cheng, 1990) and other environments, including plants, animals, and humans (Mackay, 2001). Some chemicals inherently favor and thus will migrate, when the opportunity arises, to water. These are primarily

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0.00 2.0 −0.5 k1 = −0.23 day −1 t1/2 = 3.0 days In c/c0

c, mg/L

1.5 −1.0

1.0 −1.5 0.5 −2.0 (a)

0

0

5

10 Time, Days

15

(b)

0

5

10 Time, Days

15

Figure 57.1  Dissipation rate of molinate from a rice field at 26°C as (a) a dissipation curve and (b) as a first-order plot. C0 is the initial concentration and C is the concentration at time t (see Soderquist et al., 1977 for original data; Crosby, 1998).

ENVIRONMENTAL COMPARTMENTS

TRANSFORMATION PROCESSES

Air

Volatilization

Oxidation

Soil

Sorption

Reduction

Water

Diffusion

Hydrolysis

Biota

Partitioning

Conjugation

FATE

N

ut

rie

nt

gh Su

n

Li

e ur st oi M

t ea H

d in W

ze Si

s

t

CHEMICAL

TRANSFER PROCESSES

Figure 57.2  A schematic diagram of the components of the fate of a chemical in the environment (Seiber, 1985).

chemicals of high water solubility and high stability in water, such as salts of carboxylic acid herbicides [2,4-dichlorophenoxyacetic acid (2,4-D), 4-chloro-2-­methylphenoxyacetic acid, and trichloroacetic acid]. Others favor the soil or sediment compartment because they are preferentially sorbed to soil and they may lack other characteristics (volatility or water solubility) that favor removal from soil. Examples include paraquat, which is strongly sorbed to the clay mineral fraction of soil, and DDT, toxaphene, and the cyclodiene pesticides, which sorb to and are stabilized in soil organic matter. Others, particularly fat-soluble substances, favor storage in fatty animal tissue when the opportunity arises. Organochlorine compounds including DDT, dioxins, and PCBs are examples. Volatile chemicals such as the fumigants methyl bromide and telone (1,3-dichloropropene)

and chemicals with high Henry’s law constants favor the air compartment. The elements of environmental fate prediction, based on properties of the chemical of interest that may be measured or estimated based upon structures, become apparent through these well-­established benchmark chemicals (Boethling and Mackey, 2000; Lyman et al., 1990; Mackay et al., 1992).

57.2.3  Structure The key to how a chemical will behave is contained in the chemical’s molecular makeup or structure. The field of structure–activity relationships has long been important for identifying and evaluating the spectrum of bioactivity of candidate pesticides, and now shows promise for predicting

Chapter | 57  Environmental Transport and Fate

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PROPERTIES PHYSICAL CHEMICAL khyd MW kphoto H koxidn BCF kmetab Sol. kmicro V.P. Kd

Windspeed Sunlight Intensity Humidity

Air

Flow In Temp pH Susp.Sed.

Biomass Dissolved O2

Water

Flow Out Area.

Depth

Sediment

Figure 57.3  Intrinsic and extrinsic properties governing the distribution and fate of a chemical in a pond environment (Seiber, 1987).

Table 57.1  Influence of Structure on Biological Activity, Environmental Behavior, and Regulatory Status of DDT and Dicofol

Activity as pesticide

Insecticide

Acaricide

Environmental reactivity

Stable; breakdown products (DDE and DDD) also stable

Breaks down; primary breakdown product (DCBP) also unstable

Bioconcentration potential

High in aquatic and terrestrial food chains

Low

Regulatory status (U.S.)

Banned

Still registered

environmental persistence, transport, and fate as well. (Crosby, 1998). An example of the importance of even small structural changes is provided by contrasting the behavior of the two closely related chemicals DDT and dicofol (Table 57.1). The subtle structural change, by substituting OH for H at the central carbon, has major environmental fate ramifications (as well as a strong influence on biological activity). DDT degrades slowly in the environment, and its primary breakdown products DDE and DDD are also very stable. Dicofol degrades rather rapidly in the environment, and its principal breakdown product, dichlorobenzophenone (DCBP), is also degraded further rather rapidly. DDT (and DDE/DDD) is highly lipophilic, showing strong tendencies to bioconcentrate in aquatic organisms and, through food chain accumulation, in terrestrial animals and humans. Dicofol has lower lipophilicity and more ready breakdown, both owing to the presence of the hydroxy substituents, and does not significantly bioconcentrate or bioaccumulate. Its primary breakdown products do not exhibit these negative characteristics either. Even though there has been much experience with both DDT and dicofol, new ­ information

continues to surface for the parent chemicals as well as for the degradation products. This situation is complicated because DDT can degrade by oxidation to dicofol (Crosby, 1998), and commercial dicofol may contain DDT and DDE as impurities (Qiu et al., 2005). Because of these differences in toxicity and environmental behavior, DDT was banned for use in the United States in the 1970s and dicofol is still registered for use. Thus the importance of subtle structural features cannot be over­emphasized for closely related structures such as DDT and dicofol and certainly so for more structurally diverse chemicals. If methylchlor and/or methiochlor, which are good insecticides but biodegrade rapidly in the environment (Metcalf, 1977), had been developed rather than DDT, we might still be using DDT-like insecticides in U.S. agriculture today. There has been much new study of chiral, or optically active, pesticides in recent years showing the differences in environmental behavior, particularly in microbial breakdown and metabolism and in bioactivity between enantiomers. Qin et al.’s (2008) review found that enantioselective degradation occurs frequently for pyrethroids in both field

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and laboratory. Enantiomerization, which may contribute to the enantioselectivity in pyrethroid fate and effects, is common when the pyrethroid contains an -cyano carbon and is exposed to alkalis and other conditions. The fate and effects of individual stereoisomers should be considered when predicting pyrethroid environmental impacts (Qin et al., 2008; Wong, 2006). Bidleman et al. showed that enantioselective metabolism of organochlorine pesticides can affect apparent air–soil and air–water partitioning. Analysis of enantiomers can materially aid in interpreting global transport and fate processes (Bidleman et al., 2003).

57.2.4  Activation–Deactivation Environmental transformations generally lead to products that are less of a threat to biota and the environment than the parent chemicals; that is, they result in deactivation of the parent. The products may be less toxic than the parent or have lower mobility and persistence relative to the parent. They may, in short, be simply transient intermediates on the path to complete breakdown or “mineralization” of the parent. Thus, 2,4-D may degrade to oxalic acid and 2,4-dichlorophenol. The latter is of some concern, but it lacks the herbicidal activity of 2,4-D and appears to be ­ further degraded in most environments by sunlight, microbes, etc. Organophosphates can be hydrolyzed in the environment to phosphoric or thiophosphoric acid derivatives and a substituted phenol or alcohol. These products, in the case of most organophosphates, are much less of a threat to humans and the environment than the parent chemicals. Environmental activation represents the transformations that lead to products that are more of a threat due to one or more of the following characteristics: Enhanced toxicity to target and/or nontarget organisms l Enhanced stability, leading to greater persistence l Enhanced mobility, leading to greater potential for contamination of groundwater or other sensitive environmental media l Enhanced lipophilicity, leading to bioconcentration and bioaccumulation l

Examples of activations (Coats, 1991; Wolfe and Seiber, 1993) include the formation of DDE, which is apparently the form most responsible for causing thin eggshells in birds that have accumulated DDT or DDE from their prey, and DDD, which can persist for years in some soil and water systems; formation of dieldrin and eventually photodieldrin from aldrin; oxidation of organophosphate thions to the more toxic “oxon” form; S-oxidation of aldicarb (and some other N-methylcarbamates) to the more water-soluble and, in some cases, more persistent sulfoxide and sulfone forms; formation of the volatile fumigant methylisothiocyanate (MITC) from metam sodium, the commercial precursor of MITC, when the parent is applied to moist soil; and formation of ethylenethiourea,

Hayes’ Handbook of Pesticide Toxicology

a ­ carcinogen, from ethylenebisdithiocarbamate fungicides (Wolfe and Seiber, 1993). In part because of the concern over environmental activation, the U.S. EPA requires extensive information on the occurrence, toxicity, and fate of transformation products of candidate pesticides submitted for registration (Kovacs, 1983; U.S. EPA, 2008). The tests must include significant products of hydrolysis, photolysis, oxidation, and microbial metabolism, in both laboratory and field tests, but increasingly, regulations are also geared to products that might be formed during illegal use or during fires, explosions, spills, disinfection, and other situations that expose chemicals to conditions for which they were not intended (Bourke et al., 1992). Unfortunately, not all such situations can be anticipated, requiring continual vigilance by the registrant and regulatory agencies as a part of product stewardship and environmental protection.

57.3  Environmental transport   and fate modeling Because of the complexity in the environment, and even in parts of the environment such as a pond, a cross-section of soil, or vegetation in contact with the atmosphere, environmental scientists have developed models for studying and for predicting or estimating transport and fate processes. There are two broad classes of models, described in the following sections.

57.3.1  Physical Models In a physical model, part of the environment is mimicked in a chamber that has the major features of the environment to be mimicked, but is amenable to artificial control of variables (temperature, sunlight, humidity, air flow, turbulence, etc.) in a way not possible in the real environment. A chemical of interest, or mixture, is introduced to the chamber in a way that mimics entry to the environment (e.g., in simulated rainfall, or applied to vegetation or soil). The distribution of the chemical between phases in the chamber is monitored by periodic sampling and analysis, the dissipation rate in specific phases is determined, and breakdown products are measured over a period of time. A classic chamber, one of the first constructed and tested with pesticides, was the model ecosystem of Metcalf and co-workers (1977), in which both dissipation and bioaccumulation processes were simulated for a number of pesticides and other contaminants. Even though simplified in relation to the real environment, constructing, maintaining, replenishing, and sampling such model chambers can be time-consuming and difficult. Replication is often needed to increase the precision and accuracy of the results – further multiplying the cost of operation. In

Chapter | 57  Environmental Transport and Fate

spite of these ­ challenges, physical models have received much recent attention. Vera et al. (2007) described the use of large Teflon chambers in the EUPHORF (European Photoreactor) facility to study the atmospheric fate of pesticides. Silburn and Kennedy (2007) described a rain simulation study to estimate pesticide transport in runoff. Vereecken (2005) described the mobility and leaching behavior of glyphosate in column, lysimeter, and fieldscale experiments.

57.3.2  Mathematical or Computer Models More and more environmental scientists are relying upon mathematical computer-based models in which the principle transport and fate processes are simulated using sets of equations. A “virtual” dose of pesticide is applied, and input data include both the physical and chemical properties of the pesticide as well as the size, scope, and properties of environmental variables (temperature, sunlight, moisture, etc.). The model predicts the concentration of the pesticide and its breakdown products in various compartments (air, water, soil, vegetation) at different times. Although itself only a simulation or estimation, the validity of the computer model’s output can be verified by comparing its results with a companion study in the real environment or in a chamber simulating the environment. Such models are achieving considerable status in regulatory work as a way of rapidly assessing a pesticide’s potential behavior to determine whether there might be undue persistence, bioaccumulation, activation, leaching, or some other undesirable characteristic associated with a given chemical in a specific use scenario. Recent examples include the multimedia environmental models based upon fugacity (Mackay, 2001; Mackay and Webster, 2007). Variants of this system are used as screening tools by regulatory agencies such as the U.S. EPA and California’s Department of Pesticide Regulation. Simunek et al. (2008) described a model (HYDRUS) for simulating pesticide behavior in water moving through the vadose zone. Several models were used by Inao et al. (2008) for simulating pesticide fate and transport in rice paddy environments for assessing ecological risk and its management. Other recent examples of development and use of computer models are given by Sood and Bhagat (2005), Jantunen et al. (2005), Malone et al. (2004), Dubus et al. (2003), Hansen (2001), and De Leeuw et al. (2000). Both physical and mathematical (computer) models clearly have a role in helping to describe and understand the fate of pesticides and pinpoint vulnerabilities that might exist in reaching regulatory decisions to use or restrict use of pesticide chemicals. Models are also important tools for helping to interpret monitoring data and even to direct where monitoring might be most useful in, for example, determining exposures and/or mitigation efforts.

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Conclusion Significant advances have been made in defining and understanding the principles that underlie pesticide dissipation in the environment. This has led to better testing procedures and better environmental fate models, which in turn have given rise to better methods for predicting envir­ onmental behavior or fate before use or release occurs. The development of a predictive capability has helped focus efforts in industry and agencies toward marketing and regulating safer chemicals and restricting or eliminating those that are likely to pose significant risks, in many cases early in the development stage and before widespread use occurs. As a result of this and parallel developments in understanding the mechanisms of toxic action, pesticides and pest control practices at the beginning of the 21st century are considerably safer than in the past 60 years of the synthetic pesticide era. These accomplishments provide greater protection to users of pesticides, to nearby residents, to consumers, and to wildlife.

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De Leeuw, F. A. A. M., Van Pul, W. A. J., Van Den Berg, F., and Gilbert, A. J. (2000). The use of atmospheric dispersion models in risk assessment decision support systems for pesticides. Environ. Monit. Assess. 62(2), 133–145. Dixon, R. A., Gang, D. R., Adrian J. Charlton, A. J., Oliver Fiehn, O., Harry A. Kuiper, H. A., Tracey L. Reynolds, T. L., Ronald S. Tjeerdema, R. S., Elizabeth H. Jeffery, E. H., German, J. B., Ridley, W. R., and Seiber, J. N. (2006). Applications of Metabolomics in Agriculture. J. Agric. Food Chem. 54(24), 8984–8994. Dubus, I. G., Brown, C. D., and Beulke, S. (2003). Sources of uncertainty in pesticide fate modelling. Sci. Total Environ. 317(1–3), 53–72. Fong, W. G., Moye, H. A., Seiber, J. N., and Toth, J. P. (1999). “Pesticide Residues in Foods: Methods, Techniques, and Regulations,” Wiley, New York. Food Quality Protection Act. (1996). U.S. Congress, Washington, DC. Goring, C. I., and Hamaker, J. N. (1972). Organic Chemicals in the Soil Environment. Vols. 1 and 2. Dekker, New York. Hansen, C. D. (2001) Danish EPA use of models for assessment of pesticides mobility. In “NATO Science Series, IV: Earth and Environmental Sciences, Vol. 2 (Modelling of Environmental Chemical Exposure and Risk),” pp. 183–192. Haque, R., and Freed, V. H. (ed.). (1974). “Environmental Dynamics of Pesticides,” Plenum, New York. Holland, J. (1999). Environmental fate: A down under perspective. In “Pesticide Chemistry and Bioscience. The Food-Environment Challenge” (G. T. Brooks and T. R. Roberts, eds.), Royal Society of Chemistry, Cambridge, UK. Inao, K., Watanabe, H., Karpouzas, D. G., and Capri, E. (2008). Simulation models of pesticide fate and transport in paddy environment for ecological risk assessment and management. JARQ 42(1), 13–21. Inter-Organization Programme for the Sound Management of Chemicals. (1997). “International Workshop on Endocrine Disruptors,” Report, UNEP Chemicals, Geneva. Jantunen, A. P. K., Trevisan, M., and Capri, E. (2005). Computer models for characterizing the fate of chemicals in soil: pesticide leaching models and their practical applications. In “Soil-Water-Solute Process Characterization” (J. Alvarez-Benedi and R. Munoz-Carpena, eds.), CRC Press, Boca Raton, FL pp. 715–756. Jobling, S., Reynolds, T., White, R., Parker, M. G., and Sumpter, J. P. (1995). A variety of environmentally persistent chemicals, including some phthalate plasticizers, are weakly estrogenic. Environ. Health Perspect. 103, 582–587. Johnson, S. L., and Bailey, J. E. (1999). Pesticide risk management and the United States Food Quality Protection Act of 1996. In “Pesticide Chemistry and Bioscience. The Food-Environment Challenge” (G. T. Brooks and T. R. Roberts, eds.), Royal Society of Chemistry, Cambridge, UK. Kovacs, M. F. Jr. (1983). EPA guidelines on environmental fate. Residue Rev 85, 3–16. Krieger, R. I., and Woodrow, J. E. (2007). Assessing exposure to agricultural fumigants in outdoor and indoor air environments. In “Assessing Exposures and Reducing Risks to People from the Use of Pesticides” (R. I. Krieger, N. Ragsdale, and J. N. Seiber, eds.), ACS Symposium Series 951. American Chemical Society, Washington, DC, pp. 70–86. Lehotay, S. J. (2006). Quick, easy, cheap, effective, rugged, and safe approach for determining pesticide residues. Methods in Biotechnology 19, 239–261.

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Lyman, W. J., Reehl, W. F., and Rosenblatt, D. H. (1990). “Handbook of Chemical Property Estimation Methods,” American Chemical Society, Washington, DC. Mackay, D. (2001). “Multimedia Environmental Models: The Fugacity Approach,” 2nd ed. Lewis Publishers/CRC Press, Boca Raton, FL. Mackay, D., Shiu, W. Y., and Ma, K. C. (1992). “Illustrated Handbook of Physical–Chemical Properties and Environmental Fate for Organic Chemicals,” Lewis, Boca Raton, FL. Mackay, D., and Webster, E. (2007) Simple fugacity models of off-site exposure to agrochemicals. In “Rational Environmental Management of Agrochemicals” (I. R. Kennedy, ed.), ACS Symposium Series 966, pp. 14–36. American Chemical Society, Washington, DC. Malone, R. W., Ahuja, L. R., Ma, L., Wauchope, R. D., Ma, Q., and Rojas, K. W. (2004). Application of the root zone water quality model (RZWQM) to pesticide fate and transport: An overview. Pest Manag. Sci. 60(3), 205–221. Metcalf, R. (1977). Model ecosystem studies of bioconcentration and biodegradation of pesticides. In “Pesticides in Aquatic Environments” (M. A. Q. Khan, ed.), pp. 127–144. Plenum, New York. National Research Council. (1983). “Risk Assessment in the Federal Government: Managing the Process,” National Academy Press, Washington, DC. National Research Council (1993). “Pesticides in the Diets of Infants and Children,” National Academy Press, Washington, DC. National Research Council (1994). “Science and Judgment in Risk Assessment,” National Academy Press, Washington, DC. National Research Council (1999). “Hormonally Active Agents in the Environment,” National Academy Press, Washington, DC. Organisation for Economic Co-operation and Development (OECD) (2009). Environment Directorate. Pesticides Testing and Assessment. http://www.oecd.org/document/10/0,3343,en_2649_34383_ 31951370_1_1_1_1,00.html, as of May 22, 2009. Qin, S., Liu, W., and Gan, J. (2008). Chiral selectivity in the environmental fate of pyrethroids. In “Synthetic Pyrethroids,” ACS Symposium Series 991. American Chemical Society, Washington, DC, pp. 238–253. Qiu, X. H., Zhu, T., Yao, B., Hu, J. X., and Hu, S. W. (2005). Contribution of dicofol to the current DDT pollution in China. Environ. Sci. Technol. 39(12), 4385–4390. Seiber, J. N. (1985). General principles governing the fate of chemicals in the environment. In “Agricultural Chemicals of the Future” (J. L. Hilton, ed.), pp. 389–402, Beltsville Symposia in Agricultural Research No. 8. Rowan and Allanheld, Totowa, NJ. Seiber, J. N. (1987). Principles governing environmental mobility and fate. In “Pesticides: Minimizing the Risks” (N. N. Ragsdale and R. J. Kuhr, eds.), ACS Symposium Series 336, pp. 88–105. American Chemical Society, Washington, DC. Seiber, J. N. (2005). Evolution of residue analysis and its role in improving the safety of agrochemicals. In “Environmental Fate and Safety Management of Agrochemicals” (J. M. Clark and H. Ohkawa, eds.), ACS Symposium Series 899, Am. Chem. Soc., Washington DC, pp. 14–27. Silburn, D. M., and Kennedy, I. R. (2007). Rain simulation to estimate pesticide transport in runoff. In “Rational Environmental Management of Agrochemicals” (I. R. Kennedy, ed.), ACS Symposium Series 966, pp. 120–135. American Chemical Society, Washington, DC. Simunek, J., Kohne, J. M., Kodesova, R., and Sejna, M. (2008). Simulating nonequilibrium movement of water, solutes and particles using HYDRUS-a review of recent applications. Soil Water Res. v. 3(Spec. Iss. 1), S42–S51.

Chapter | 57  Environmental Transport and Fate

Soderquist, C. J., Bowers, J. B., and Crosby, D. G. (1977). Dissipation of molinate in a rice field. J. Agric. Food Chem. 25, 940–946. Sood, C., and Bhagat, R. M. (2005). Interfacing geographical information systems and pesticide models. Curr. Sci. 89(8), 1362–1370. Suter, G. W., Barnthouse, L. W., Bartell, S. M., Mill, T., Mackay, D., and Paterson, S. (1993). “Ecological Risk Assessment.,” Lewis, Boca Raton, FL. Thibodeaux, L. J. (1979). “Chemodynamics: Environmental Movement of Chemicals in Air, Water, and Soil,” Wiley – Interscience, New York. Thomas, B. (1991). Pesticide registration in Europe. In “Regulation of Agrochemicals” (G. J. Marco, R. M. Hollingworth, and J. R. Plimmer, eds.), pp. 73–79. American Chemical Society, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA). (1982). “Pesticide Assessment Guidelines. Subdivision N. Chemistry: Environmental Fate.” EPA-540/9-82-021, Office of Pesticides and Toxic Substances, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA) (2008). Data Requirements for Pesticides. Subpart N—Environmental fate. Code of Federal Regulations, Title 40, Part 158.1300. U.S. Environmental Protection Agency (U.S. EPA) (2009a). Test guidelines for data requirements. http://www.epa.gov/pesticides/science/ guidelines.htm, dated April 20, 2009. U.S. Environmental Protection Agency (U.S. EPA) (2009b). OPPT Harmonized test guidelines, Series 835: Fate, Transport and

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Transformation Test Guidelines. http://www.epa.gov/opptsfrs/publications/OPPTS_Harmonized/835_Fate_Transport_and_Transformation_ Test_Guidelines/index835.html, Dated April 15, 2009. Van Emon, J. M. (2007). “Immunoassay and Other Bioanalytical Techniques,” CRC Press, Boca Raton, FL. Vera, T., Munoz, A., Mellouki, A., Rodenas, M., and Vazquez, M. (2007). The use of EUPHORE facility for studying the atmospheric fate of pesticides. In “Environmental Fate and Ecological Effects of Pesticides,” Symposium Pesticide Chemistry, 13th, Piacenza, Italy, Sept. 3-6, 2007 (A.A.M. Del Re, ed.) pp. 278–285. Vereecken, H. (2005). Mobility and leaching of glyphosate: a review. Pest Manag. Sci. 61(12), 1139–1151. Wolfe, M. F., and Seiber, J. N. (1993). Environmental activation of pesticides. In “De Novo Toxicants: Combustion Toxicology, Mixing Incompatibles, and Environmental Activation of Toxic Agents” (D. J. Shusterman and J. E. Peterson, eds.), Occupation Medicine: State of the Art Reviews, Vol. 8, pp. 561–573. Hanley and Belfus, Philadelphia. Wong, C. S. (2006). Environmental fate processes and biochemical transformations of chiral emerging organic pollutants. Analyt. Bioanalyt. Chem. 386(3), 544–558. Younglai, E. V., Wu, Y. J., and Foster, W. G. (2006). Do insecticides have adverse effects on reproduction. Immunol. Endocr. Metabol. Agents Med. Chem. 6(1), 45–56.

Chapter 58

Hydrophobicity as a Key Physicochemical Parameter of Environmental Toxicology of Pesticides Toshio Fujita1, Keiichiro Nishimura2, Chiyozo Takayama3, Masanori Yoshida and Matazaemon Uchida4 1

EMIL Project; 2Osaka Prefecture University; 3Sumitomo Chemical Company; 4Nihon Nohyaku Company

58.1  Introduction The environmental behaviors of pesticides, such as accumulation in soil, contamination of aquasphere, residue levels in crops, and bioaccumulation through food chains as well as nondietary routes, are dependent on their distribution properties among various environmental phases. These distribution features are modeled by phase-distribution equilibrium constants such as soil absorption coefficient, water solubility, and bio-concentration factors in biota, including crops (Briggs, 1981a). As early as 3 decades ago, the logarithm of these constants was recognized to be related with a reference parameter representing the molecular “hydrophobicity” (Briggs, 1969, 1981a; Hance, 1967; Hansch et al., 1968; Neely et al., 1974; Valvani and Yalkowsky, 1980). The most frequently used hydrophobicity parameter is the log P [or log k(o/w)], P [or k(o/w)] being the 1-octanol/ water partition coefficient (Fujita et al., 1964; Leo, 1993; Noble, 1993). Moreover, the log P value of organic compounds has been shown to be the most decisive parameter for their toxicity brought about by nonspecific perturbation of biomembraneous and enzyme systems (Hansch and Leo, 1995). Specific toxicities to target pests are governed by specific mechanisms defined by various physicochemical properties of the molecule, including the hydrophobicity (Hansch and Leo, 1995). Thus, the environmental toxicology of pesticides covering distribution patterns, persistence, and toxicity could be quantitatively analyzable in terms of physicochemical molecular descriptors including hydrophobic, electronic, steric, and others with the use of regression analyses, in which the log P value is regarded as playing a central role (Hansch and Leo, 1995). That is, the QSAR (quantitative structure– activity relationship) procedure initiated and developed by Hayes’ Handbook of Pesticide Toxicology Copyright © 2001 Elsevier Inc. All rights reserved

Hansch and co-workers (Hansch and Fujita, 1964, 1995; Hansch and Leo, 1995) could apply to various environmental aspects of pesticides in spite of the tremendous complexity of the processes involved. In this chapter, we review the measurement and estimation procedures of the log P for a wide range of organic compounds and the significance of the log P value in elucidating and predicting the environmental behavior of pesticides in terms of the QSAR.

58.2  Measurements and experimental estimations of log p There have been a number of experimental procedures either to measure directly or to estimate indirectly the log P value of various types of compounds (Sangster, 1997a). Among them, the shake-flask and slow-stirring procedures as direct methods and the potentiometric method for ionizable compounds and the high-performance liquid chromatographic procedure as an indirect method are described in this section. The shake-flask and liquid chromatographic procedures are now standardized by the Organization for Economic Co-operation and Development (OECD) (1981, 1989) for the assessment of the environmental effects of organic compounds, including pesticides.

58.2.1  Direct Partitioning (Shake-Flask and Slow-Stirring) Method In the measurements of the P value in the 1-octanol/water system, the ratio of concentrations in the two phases should be calculated after the partitioning equilibrium is established as far as possible (Fujita et al., 1964). Analytically pure 1-octanol is commercially available. Impurities, if any, can 1229

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be removed by consecutive washings with 2 N H2SO4, 2 N NaOH, and water. Sufficiently pure 1-octanol is obtained by distillation after desiccation. Distilled water, under decarbonated or buffer conditions if necessary, should be used for the partitioning. The two solvent phases should be mutually saturated before the partitioning (Smith et al., 1975). In most cases except for highly hydrophobic and hydrophilic solutes, the equilibrated concentration is measured only in the water phase. The concentration in the 1-octanol phase is calculated after subtraction of the amount in the water phase from the total. The partitioning experiment is usually repeated several times, varying the amount of solute and volume ratios of 1-octanol and water. To minimize the calculation error, the volume ratio should be chosen so that the amount of the solute in the two phases after equilibrium is equal, or nearly so. Depending on its polarity, the test compound is dissolved in either the pretreated octanol or the pretreated water phase. The amount is selected so that the initial concentration is on the order of 103 M or less (to discourage dimer or micelle formation), but sufficient so that the equilibrated concentration in the water phase can be accurately measured directly, whenever possible (i.e., without concentration via reextraction or evaporation). For moderately functionalized and polar compounds, the log P of which is neither too high nor too low, a consistent log P value can sometimes be obtained by 100 inversions of the flask in 5 min or less (Leo et al., 1971). However, if the solute does not have surfactant properties, “vigorous” agitation for 1 h seems to be better, in general, for hydrophobic pesticides. If the agitation is mild, it takes the partition equilibrium a longer time to be achieved. After the partitioning agitation, the two phases are usually separated with centrifugation at 2000–3000 rpm for 15 min. Note that a shorter time at a higher rpm can still leave the aqueous phase supersaturated with 1-octanol. In cases in which the partitioning procedure yields emulsion in both phases (or between two turbid solvent layers even after centrifugation), filtration of the aqueous phase through a minimum amount of “inert filtration aids” under a slightly reduced pressure is effective to make the phases clear. Celite powders (Highflo-super-cel, Ø: 7 m, Celite Corp., Lompoc, CA) are conveniently used after being packed tightly with a minimum amount of the turbid octanol on a cotton ball placed on a funnel neck. The adsorption of solutes from the water phase to filtration aids has been shown to be negligible as far as compounds with a log P value between 1 and 2.5 are concerned (Nishikawa, 1989). Although the P values are not affected much by temperature variations (Fujita et al., 1964; Leo et al., 1971), it is preferable to keep the temperature within a few degrees of 25°C during the entire operation (Dearden and Bresnen, 1988). Vigorous agitation with a “complete” mixing of the two phases is not applicable to highly hydrophobic compounds

Hayes’ Handbook of Pesticide Toxicology

with a log P value greater than 4–5 because of a persistent emulsification of the system. The emulsified phases are difficult to clear even with a prolonged centrifugation. The celite filtration would not be favorable in this case because a stronger adsorption of compounds from the aqueous phase is probable. For such compounds, usually with a high molecular weight and only poorly functionalized, the slow-stirring procedure is recommended. In this procedure, the two phases are equilibrated under conditions of slow stirring (about 200 rpm for 1000 ml of the system, including 20–50 ml octanol phase) (Brooke et al., 1986; de Bruijn and Hermens, 1990; de Bruijn et al., 1989). It takes much time for the equilibrium to be established because any mixing between the two phases should be avoided as far as possible. For compounds with a log P value lower than 4, the equilibrium occurrs within 1–2 days, but, for compounds of log P 5, it takes 2–3 days. For p,p-DDT of which log P value was expected to be higher than 6 (6.29, as estimated from that of methoxychlor, 4.83; Nishimura and Fujita, 1983), 3–4 days of stirring were required with a resultant log P of 6.91 (de Bruijn et al., 1989). The reproducibility of the procedure is reported to be very good, the standard deviation among five measurements being 0.03 for p,p-DDT (de Bruijn et al., 1989). For highly hydrophobic compounds of log P 5, the amount of solute in the octanol phase is only negligibly decreased (0.1%) at the equilibrium from the total in the initial state with a volume ratio of 1:100 (octanol/H2O). If only the water phase is analyzed, it is very important to measure the very low concentration with a high accuracy. For nonvolatile solutes, this could be done by extracting the solute from the water phase with hexane followed by evaporation to give a concentration range in which the solute can be analyzed with a sufficient precision. For highly hydrophilic compounds with log P  5, the reverse situation occurrs. A very low concentration in the octanol phase must be measured accurately. Condensation of the octanol phase under a reduced pressure and an appropriate dilution with methanol or acetone-hexane (1:1, v/v) may make it possible to analyze the amount of solute fairly precisely. An example for the highly hydrophilic glyphosate is published recently (Chamberlain et al., 1996). To measure the log P value of ionizable compounds, 0.1–0.01 N HCl and NaOH solutions are used as the aqueous phase for acidic and basic solutes, respectively. Buffer solutions can also be used under conditions in which their buffer capacity should be enough to make the solutes in the aqueous phase exist entirely as the nonionized neutral form. The buffer should not be extractable into the octanol phase so that phosphate buffers are best recommended (Wang and Lien, 1980). The log P value of organic ion pairs, in which either or both the cationic and the anionic species are organic, is sensitive to variations

Chapter | 58  Hydrophobicity as a Key Physicochemical Parameter of Environmental Toxicology of Pesticides

in the ionic strength of the aqueous phase (Takayama et al., 1985; Terada et al., 1981). The partition behavior of multiprotic compounds such as peptides, including insect neurohormones (Menn and Borkovec, 1989), is also modified by (inorganic) counterions as well as ionic strength in the aqueous phase (Akamatsu et al., 1989). Therefore, experimental conditions for the measurement of the “true” log P value of noncharged species of ionizable pesticides should be carefully controlled. It should be mentioned that charged molecular species or ions have no unique hydrophobic index in terms of log P. It varies depending on the counterionic species as well as ionic strength in the aqueous phase. For gaseous compounds used as fumigants and spraypropellants, as well as environmental contaminants such as halomethanes, fluoroalkanes, and nitrogen oxides, specially designed partitioning apparata and gas chromatographic systems are required to measure the accurate log P value (Hansch et al., 1975). The concentration or the amount of solute in the water phase and, if necessary, also in the octanol phase is analyzed by the appropriate spectrometric, chromatographic, radiometric, and gravimetric (Masutani et al., 1981) procedures. This analysis is performed directly after centrifugation or with an appropriate combination of extraction, condensation, and solvent evaporation, depending on the situation. The ultraviolet spectrometric analysis is most widely used, particularly for aromatic compounds. From the spectra, it is possible to confirm whether or not the structure of the solute molecule is transformed during the partitioning process (Fujita et al., 1964). The gas (Sotomatsu et al., 1987) and high-performance liquid chromatographic analyses (Akamatsu et al., 1989) are also convenient for aliphatic as well as volatile aromatic compounds. In these cases, the confirmation can be made by comparison of retention time before and after partitioning. It is the most important prerequisite in the log P measurement that no structural change occurs during the partitioning procedure. Sometimes, the procedure should be done under nitrogen atmosphere or with an addition of a minimum amount of sodium thiosulfate to the aqueous phase to prevent air oxidation (Fujita et al., 1964). Radiometric analysis are sometimes used for solutes at very low concentration in the aqueous or octanol phase, that is, highly hydrophobic in the water (Uchida et al., 1974) or highly hydrophilic in the octanol (Chamberlain et al., 1996). Serious error can occur in this procedure if radiocolloids adsorb on the inner wall of the vessels. This is in addition to any errors introduced from lack of purity of the labeled compounds. To reduce the effect of the radiocolloid adsorption, an appropriate amount of the corresponding nonlabeled compounds (carriers) should be added to the system to lower the radiocolloid formation by the chemical dilution of the radioactivity (Keller, 1993).

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58.2.2  Potentiometric Titration Method for Ionizable Pesticides A potentiometric titration method for measuring the pKa of ionizable pesticides has been reported, which gives log P values at the same time (Chamberlain et al., 1996). The method typically involves two titrations: The first is to measure the pKa value in the water phase and the other ′ is for an apparent “pKa” value, pKa , after addition of an appropriate volume of 1-octanol. The difference, pKa   pKa′   pKa(0), is a function of the P value of the neutral form as shown in the following equation, where V is the volume of the solvent phase (Kaufman et al., 1975):

P  VH2 O [log1 pKa  1]/Voct .



(1)

The entire operation of titration and calculation of pKa and log P values has been automated (Avdeef, 1991; Clarke, 1984). It has been developed for measurement of stepwise pKa values as well as the true log P values of noncharged species and apparent log P values of ion-paired counterparts for multiprotic compounds (Avdeef, 1992). The true log P value more negative than 2 for neutral species cannot be measured using this method because the pKa value is too small to estimate accurately. Also the volume ratio of 1-octanol/water should be high, making it difficult to titrate the octanol/water system including the test compound. The true log P value of amitrole (0.97) was measurable, but that of glyphosate (3.39 by the shake-flask procedure) was not (Chamberlain et al., 1996).

58.2.3  High-Performance Liquid Chromatographic Method There are review articles dealing with reversed-phase highperformance liquid chromatography (RP-HPLC or HPLC) applied to the experimental estimation of log P values (Braumann, 1986; Terada, 1986). In HPLC, the affinity of a solute for the stationary phase is characterized by the retention factor k, which is defined as

k  ( t R  t 0 ) /t 0 .

(2)

where tR is the retention time of the solute and t0 is that of a nonretained reference compound. Because the chromatographic retention can be regarded as an equilibrium partition process between two “immiscible” phases, the log P of solutes in the 1-octanol/water system is regarded as being linearly related to the log k empirically as

log P  a log k  b.

(3)

The coefficients a and b in Eq. 3 are calculated by the regression analysis from the log k values measured for a series of related compounds whose log P values have been

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1232

determined by the shake-flask procedure. If the statistical quality of Eq. 3 is sufficiently high, then the log P values of other compounds belonging to the same series can be estimated from their log k values and Eq. 3. It is unfortunate, however, that the stationary phase is not entirely able to simulate the octanol phase. The mobile-phase eluent is a solvent/water (solvent/buffer) mixture of certain proportions, the most commonly used solvent being methanol. Thus, the retention (partition) behavior of compounds in the HPLC system differs from that in the 1-octanol/water system according to physicochemical differences of each of the two phases between the two experimental systems. As the stationary phase, various types of ODS (octadecyl silica trapping some amount of organic solvent from the mobile phase) columns have been most widely used. According to the difference in the hydrogen-bonding characteristics with the 1-octanol/water system, the solutes are generally grouped into non-hydrogen bonders, hydrogen acceptors, and amphiprotics. For each of the groups, a separate linear relationship of the type of Eq. 3 can be drawn in such a way that the a value is almost equivalent but the constant b differs among groups. Thus, hydrogen-bonding indicator variable terms (Fujita et al., 1977) for groups can be added to Eq. 3 (Terada, 1986). With an ODS column and an acetonitrile/water eluent, Takahashi et al. (1988) measured log k values for a set of fungicidal substituted N-phenylcarbamates having various substituents on the benzene ring. From a significant correlation of the type of Eq. 3, observed for 31 analogs with log P values measured by the shake-flask procedure, the log P values for 38 new compounds were estimated from their log k values and used for the QSAR studies. It has been proposed that the log k value for a given compound increases linearly, over a moderate range, with decrease in proportion of the miscible organic solvent in the eluent (Karger et al., 1976; Miyake and Terada, 1982). Recently, Yamagami et al., (1994) have observed that the linearity is not always valid. They also demonstrated that the log k0 value defined by the extrapolation of log k values to the zero organic-solvent concentration in the mobile phase is not necessarily a useful predictor of the log P value. The use of eluents containing 50% (v/v) methanol gives the most reliable and straightforward correlation between log k and log P for a series of related compounds having non-hydrogen-bonding and hydrogen-accepting substituents. For those having amphiprotic substituents, the situation is less straightforward, requiring additional parameters (Fujita et al., 1977). Yamagami and Fujita (1995) and Yamagami et al. (1990, 1994, 1995) have further indicated that the correlation of the type of Eq. 3 should generally be dealt with so as to include, besides the previously mentioned indicator variable terms, parameter terms representing mutual electronic effects among substituents or functional groups in a number of substituted heteroaromatic compound series. These electronic effects are thought to govern variations in the hydrogen-bonding

interactions of solutes with the mobile phase relative to that with the stationary phase. With improvements in the chromatographic system as well as elaborations in the combinations of solid support and mobile phase, the HPLC method is reported to work very well for estimation of the log P ranging from 0 to 6 (Sangster, 1997a). The most convenient aspect of the HPLC method is that the analytical measurements are not required. This advantage can only be guaranteed, however, with highly qualified model correlations of the type of Eq. 3.

58.3  Nonexperimental estimations of log P Experimental log P values are correctly measurable/estimable under careful conditioning. However, experimental procedures are sometimes time consuming, especially for highly hydrophobic and hydrophylic compounds. Moreover, the log P values of compounds are required before synthesis in order to assess their environmental behaviors as well as biological activity. Using recently developed methods of combinatorial synthesis, an enormous number of possible pesticide candidates are synthesized almost simultaneosly. To measure the log P value of each compounds experimentally is nearly impossible within a given period of time. Thus, nonexperimental procedures, preferably computerized, for the log P estimation are greatly needed. It should be mentioned that the construction of any estimation procedure is based largely on experimental log P values of existing compounds analyzable in terms of their physicochemical nature and composition. There is always the need for a reliable training set of standard experimental log P values to judge the performance of the estimation procedure. After an introductory section about the additive– constitutive nature of log P values, some empirical and computerized procedures will be described in this section. For other procedures, especially computerized systems, review articles (Leo, 1993; Sangster, 1997b) should be consulted.

58.3.1  Scope and Limitation of the Additive Nature of Log P Values The log P value of a molecule had been recognized to be roughly expressible by the sum of the hydrophobicity indices attributable to submolecular components such as a parent skeleton and substituents (Fujita et al., 1964; Iwasa et al., 1965). For a monosubstituted aromatic compound PhX, the hydrophobic constant  of the substituent X is defined as follows, where (X/PhH) denotes that X is to be introduced into unsubstituted benzene PhH:

(X/PhH)  log P(PhX)  log P(PhH).

(4)

Chapter | 58  Hydrophobicity as a Key Physicochemical Parameter of Environmental Toxicology of Pesticides

On certain occasions, there is a simple additivity so that the log P of disubstituted benzene PhXY can be expressed as log P(PhXY)  log P (PhH)  (X/PhH )  (Y/PhH ) (5) However, such a simple additivity is severely limited and does not hold in general. As a minimum requirement for Eq. 5, X and Y should be neither capable of hydrogen bonding nor located close together (vicinally). In certain PhXY series, X and Y usually denote variable and fixed substituents, respectively. In general, the (X/PhY) value of a certain X substituent is not a constant among various PhY systems, but varies depending on the nature of, as well as the location relative to, the Y substituent (Fujita et al., 1964). The hydrogen-bonding solvation of the Y substituent, when it is capable of hydrogen bonding, with octanol relative to that with water varies depending on the electron-withdrawing character of the X substituent. The susceptibility of variations in the relative solvation of the electronic effect of X differs among various PhY series according to the structural features of the Y substituent. The situation will be illustrated taking the X-substituted phenols (Y  OH) and the corresponding X-substituted benzoic acids (Y  COOH) as examples. The  value of a certain X substituent in these disubstituted benzenes, being defined as log P(PhXY) – log P(PhY), differs between the two systems. Besides the “intrinsic” hydrophobicity of the X substituent, the difference in the pattern of the solvation with partitioning solvents between OH and COOH is involved by definition, leading to the variations in the (X/ PhY) value (see Section 58.3.4 for further references). For the aliphatic substituents, the  value also varies depending on the stereoelectronic situations in which the substituents are located. Thus, the (X) value in the aliphatic system, (X/RH), where R is the alkyl, differs from that in the aromatic system, (X/PhH). The only exception is the  value of eletronically nearly neutral alkyl groups. (H) is always 0 by definition. Structural factors such as branching, cyclization, and multiple bonding, as well as inductive interaction between polar groups through single bonds and conjugation of -electron systems are known to contribute to variations in the molecular log P value. Table 58.1, lists the (X/PhH) and (X/RH) of some common substituents. The structural factors participating in the log P values will be described in Section 58.3.5.

measured log P value, 5.49, agrees well with the estimated value, considering the difficulties that were supposed to have occurred in the shake-flask procedure (Nakagawa et al., 1982). O

CH3OCO COOCH2

I

X

COOCH2

II

X

S

COOCH2

X

III





log P(Phenothrin, I:X  m-OPh )  log P(Methy 1R-trans -chrysanthemate) 3.76 (measured; Nakagawa et al., 1982)  log P(CH3 COOCH 2 Ph) 1.99 (measured; Nakagawa et al., 1982)  log P(CH3 COOMe) 0.18 (measured; Hansch et al., 1995)  (OPh/PhH) 2.08 ("measured"; Hansch et al., 1995)  7.65 (6)

The second and third terms in Eq. 6 represent the [Ph/(CH3COOCH2)H]. The stereoelectronic environment of the Ph group in benzyl acetate and that of the Me group in methyl acetate can be simulated well by that of the Ph group in benzyl chrysanthemate and that of the Me group in methyl chrysanthemate, respectively. The addition of the (OPh/PhH) value does not seem to be correct enough because it is from the nonsubstituted benzene system. The aromatic system, to which the OPh substituent is introduced, has, however, a substituted “toluene” (benzyl) substructure, so that the electronic effect of the OPh on the solvation of the ester (–OCO–) grouping is attenuated by the methylene unit. In fact, the [X/(CH3 COOCH2 C6H 4)H] values were experimentally shown to be almost equivalent to the corresponding (X/PhH) values (Nakagawa et al., 1982). The log P values of other substituted benzyl chrysanthemates (I) were estimated by Eq. 6 using (X/PhH) in place of (OPh/PhH) (Nakagawa et al., 1982). Similar procedures were used for substituted benzyl pyrethrates (II) (Nishimura et al., 1987) and kadethrates (III) (Matsuda et al., 1989) in which the log P value of the methyl esters of skeletal acids was experimentally measured. The log P value of an experimental amide-type Hill reaction inhibitor (Shimizu et al., 1988) was derived as follows:

58.3.2  Empirical (Manual) Procedure With the preceding limitations, one should be very careful to use the “simple” additivity principle “manually.” Some appropriate examples follow. Equation 6 is for a synthetic pyrethroid, phenothrin. The first three terms together, 5.57 ( 3.76  1.99  0.18), represent the log P value estimated for benzyl chrysanthemate (I: X  H). The experimentally

1233



log P(3, 4-Cl2 C6 H3 NHCOCH 2 OPh )  log P(3, 4-Cl2 C6 H3 NHCOCH 2 Ph ) 4.85 (measured; Mitsutake et al., 1986)  log P (PhOCH 2 CONH 2 ) 0.76 (measured; Hansch et al., 1995)  log P(PhCH 2 CH 2 CONH 2 ) 0.91 (measured; Iwasa et al., 1995)  4.70 (7)

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1234

Table 58.1  Hydrophobicity Parameter  of Common Substituents Substituent X

(X/PhH)a

(X/RH)b

Substituent X

(X/PhH)a

(X/RH)b

F

0.14

0.73

OH

0.67

1.80

Cl

0.71

0.13

OMe

0.02

0.98

Br

0.86

0.04

OEt

0.38

—c

I

1.12

0.22d

OPr

1.05

—c

Me

0.56

0.53e

OCF3

1.04

—c

Et

1.02

—c

OCONHMe

0.97

—c

Pr

1.55

—c

OCH2COOH

0.79

—c

i-Pr

1.53

—c

NH2

1.23

1.85

1.03

—

c

NMe2

0.18

0.95

—

c

NHAc

0.97

—c

0.16f

NHCONH2

1.30

—c

CH2OH CH2COOH CF3

0.72 0.88

CHO

0.65

—c

CONH2

1.49

2.28

Ac

0.55

1.26

N02

0.28

1.07g

COOMe

0.01

0.91

SMe

0.61

—c

COOH

0.32

1.26

SCF3

1.44

—c

CN

0.57

1.47

SO2Me

1.63

—c

—c

SO2NH2

1.82

—c

Ph

1.96

a

From Hansch and Leo (1995). Unless noted, R)  PhCH2CH2CH2 from Iwasa et al. (1965). The (X/RH) varies depending on the structural features of R. See Section 58.3.5. Not available, but calculable by the CLOGP procedure described in Section 58.3.5. d The log P value of PhCH2CH2CH2I is from Leo et al. (1971). e R  PhCH2CH2. f R   PhCH2CH2, taken from Takayama et al. (1985). g R  PhCH2CH2, taken from Hansch et al. (1995). b c

The second and third terms together express the  log P value between the OPh and CH2Ph compounds having the XCH2CONH2 structure, which is actually equivalent to   [OPh/H(CH2CONH2)]  [CH2Ph/H(CH2CONH2)]. This difference is considered to simulate the difference in the log P value between molecules with the XCH2CONHPI1 structure. As the previous two examples show, the stereoelectronic environments of the reference compounds or reference substituents should be carefully selected so as to be as close as possible to those of compounds the log P value of which is to be estimated with this procedure.

58.3.3  Empirical Procedure Using Relationships with Log P Values of   Simpler Compounds The following equation was derived from experimentally measured log P values for a series of N-acyl-N9-alkyl- and

N-acyl-N9-phenylureas, including insecticidal benzoylphenylureas (IV) in which X1, X2, and Y were variously changed (Sotomatsu et al., 1987). X1 CONHCONH X2



Y

IV

log P (RCONHCONHR)  0.955(0.031)[log P (RCONH 2 )  log P(NH 2 CONHR)]  1.938(0.054), n  10, s  0.072, r  0.0999.

(8) In this and the following equations, n is the number of compounds, s is the standard deviation, and r is the correlation coefficient, whenever applicable. The numbers in parentheses represent the 95% confidence intervals. R and R are either simple alkyl or (un)substituted

Chapter | 58  Hydrophobicity as a Key Physicochemical Parameter of Environmental Toxicology of Pesticides

phenyl existing in structure IV. Equation 8 indicates that the molecular log P value of the N-acyl-N’-substituted ureas can be linearly related to the sum of the log P values of the component-substituted amide and urea. The effect of the R group on the CONH2 in amides is almost linear (with a slope of nearly unity) to that on the (CONH)2 bridge in the molecule. The effect of the R group on the NHCONH2 substituent in R-substituted urea is also almost linear (with a slope of nearly unity) to that on the (NHCO)2 moiety. In this simple combination, the NH between two carbonyl groups is counted twice. The total effect of the R and R substituents on this NH group is incorporated into the summation of the log P values. The difference between the observed value and the summation term would be, however, almost constant and included in the constant term in Eq. 8. Equation 8 also shows that the (X/PhCONH2) as well as the (X/ PhNHCONH2) values can be used to estimate the log P values of substituted benzoylphenylureas with the value of the experimental log P (PhCONHCONHPh) ( 3.39) (Sotomatsu et al., 1987). The log P value of 10 dibenzoylhydrazines (V) acting as ecdysone agonists to various extents was nicely analyzed to give the following with the use of the corresponding submolecular log P value (Oikawa et al., 1994). t-Bu CONNHCO Cl





Xn

V

log P (Compounds V)  1.008(0.040) log P(X n PhCONH 2 ) 0.158(0.30) ∑ Esortho  1.952(0.046), n  10, s  0.015, r  0.999 (9)

Equation 9 clearly indicates that the log P of compounds V, in which the substituents X(n) are varied, corresponds nicely with that of the X(n)-substituted benzaortho mides. ∑ Es is the summation of the E value of the s

X substituent(s) located at the ortho position(s) as defined by Taft (1956) (for aliphatic substituents) and extended by Kutter and Hansch (1969). Note that the reference point has been shifted to that of hydrogen so that ES(H)  0. The Es value is defined so that the bulkier the substituents, ortho the more negative the value. Thus, the negative ∑ Es term means that, by increasing the sum of the size of the ortho substituents (X ortho , n  1 or 2) in compounds V, n the log P value increases relative to that of the corresponding benzamides. This equation was used to the estimate log P values of some 60 analogs ranging from 2.0 to 5.5 (Oikawa et al., 1994).

1235

58.3.4  Empirical Procedure Using FreeEnergy Related Substituent Parameters The preceding procedures can be used in a limited series of compounds under conditions where the log P values for the reference series of (simpler) compounds are either available or measured. A more general procedure is certainly needed. The (X/PhY) value of the substituents in a series of meta- and para-substituted phenols and anilines (X)  C6H4  Y:Y  OH or NH2) was shown to be related to the (X/PhH) value in monosubstituted benzenes in the following manner (Fujita et al., 1964):

  (X/PhY)  (X/PhH )  Y X  c. (10)

In Eq. 10, X is the Hammett constant of each nonhydrogen-bonding substituent, X, representing its electronic effect on the relative hydrogen-bonding solvation of the fixed substituent Y (OH or NH2). Y is the susceptibility constant of the Y substituent to the variations in the electronic effect of X. The intercept c should be as close as to 0. This type of procedure for analyzing the hydrophobicity parameters of substituted benzene systems using free-energy related substituted parameters has been continued. For meta- and para-disubstituted benzenes, X  C6H4  Y, where X and Y are both capable of hydrogen bonding, the  value in Eq. 10 depends not only on the “forward” electronic effect, X of the X substituents on the Y substituent but also on the “backward” effect of the Y substituent on the X substituents. This situation is represented in general as (Fujita, 1983; Fujita et al., 1964) 0 0   (X/PhY)  (X/PhH )  Y X  X Y  c. (11)

0 is one of the variations of the Hammett  value used in situations where any through-resonance interaction does not occur directly between the X and Y substituents (Yukawa and Tsuno, 1959). For the log P values of meta- and para-disubstituted benzenes relative to that of unsubstituted benzene, Eqs. 10 and 1110,11 are modified as follows, respectively, in which the slash within the parentheses following the  log P notation means just “relative to” (Nakagawa et al., 1992):





 log P (X  C6 H 4  Y/PhH )  ∑ (X, Y/PhH )  Y X0  c,



(12)

 log P (X  C6 H 4  Y/PhH )  ∑ (X, Y/PhH )  Y X0  X Y0  c. (13)

For meta- and para-substituted benzoic acid (X  C6H4  COOH), the preceding situations are represented as follows. For derivatives in which X is not capable

Hayes’ Handbook of Pesticide Toxicology

1236

of hydrogen bonding, Eq. 12 is enough without considering the backward effect of the COOH function on the non-hydrogen-bonding X substituents (x  0) to give (Fujita, 1983)  log P  0.964(0.071)



∑ (X, COOH/PhH)

0.499(0.140)X0 (m, p) 0.021(0.048), n  12, s  0.035, r  0.997.

(14)

In Eq. 14, the ∑  term corresponding to the second term on the left-hand side of Eq. 12 is moved to the right-hand side. Although it should be close to unity, the slope of this ∑  term is not necessarily equal to unity. As a counterpart of Eq. 13 for cases in which the set of variable X includes hydrogen-bonding substituents, the following was derived (Sotomatsu et al., 1993):  log P  0.999(0.042)



∑ (X, COOH/PhH)

(15) 0.419(0.095)X0 (m, p) 0.383(0.115)X (m, p)  0.019(0.044), n  22, s  0.049, r  0.998.

In Eqs. 13 and 15, the X value is the susceptibility parameter of the X substituents capable of hydrogen bonding to the (backward) electronic effect of the fixed Y (COOH) substituent. It is used as an independent variable. It is estimated as being equivalent to the Y value in the correlation of the type of Eq. 14 for a series of compounds in which X is now fixed and Y is varied only within nonhydrogen-bonding substituents (X is replaced by Y in Eq. 12). The X value of non-hydrogen-bonding substituents is 0. The regression coefficient of the X term corresponds to the Y0 value of the Y substituent. Thus, according to the relative position of the X and Y substituents, two x [x(m) and x(p)] terms should “theoretically” be considered. Because the Y0 (meta) and Y0 ( para ) values as the coefficients of the two x terms were close and no statistical difference was observed, the two independent terms are combined into the single X Y0 term in Eq. 15. For a set of  values, including those of the ortho-X substituents (Fujita, 1983; Fujita and Nishioka, 1976), the “regular” electronic effect of the ortho substituents is taken as being equivalent to that of the corresponding para substituents, so that Y0 (ortho)  Y0 ( para ), whereas the “proximity” electronic effect is expressible by such an inductive effect parameter as I, defined by Charton (1981). In addition, the proximity steric effect of ortho substituents is represented by the Taft–Kutter–Hansch Es parameter mentioned in Section 58.3.3 (Kutter and Hansch, 1969). Because the proximity effects also work bidirectionally, Eq. 13 can be modified as  log P( X  C6 H 4  Y/PhH)  ∑ (X, Y/PhH)

In Eq. 16, X0 , IX , EsX , X , IX , and X are used as independent variables. The first three variables are for the “forward” effect of variable X substituents on the Y substituent, whereas the next three are for the susceptibility to the “backward” effect of the Y substituent on the hydrogen-bonding X substituents. To extend the procedure toward a large number of multisubstituted compound series in which various substituents capable of hydrogen-bonding are involved, differentiation between the forward and backward effects is not straightforward. Instead, Eq. 16 can be generalized so that only the forward effects of every other substituent are considered on each of the hydrogen-bonding substituents. Because the forward effects are generally composed of three (regular and proximity electronic and steric) components, the extended correlation equation to analyze as well as to predict the  log P [ log P  log P(PhH)] value takes the form

 Y X0 (o, m, p)  X Y0 (o, m, p)  IY IX (o)  IX IY (o)   Y EsX (o)   X EsY (o)  c. (16)



 log P (Xi -benzene/benzene)  a ∑ (Xi /PhH)  ∑ [ ∑  0 (o, m, p)  I ∑ I (o)   ∑ Es (o)]  c.



(17)

∑

covers all substituents, Xi, on the benzene ring. The second term on the right-hand side takes care of the three 0 components of the “forward” effect. ∑  , ∑ I , and

∑ Es are made for substituents at positions indicated in parentheses relative to individual hydrogen-bonding sub-

stituents. The ∑ sign outside the brackets means to sum up the forward effects on every hydrogen-bonding substituent. Each of the a, , I, , and c values is calculated by regression analysis (Nakagawa et al., 1992). The preceding procedure, counting substituent effects toward every hydrogen-bonding substituent “forwardly” and “multiply,” has been nicely used to analyze some 200 log P values of multisubstituted benzenes (n  210, s  0.118, r  0.994) with some corrections for intramolecular hydrogen bonding and a buttressing effect of vicinally located substituents on the solvation of hydrogen-bonding substituents (Nakagawa et al., 1992). Recently, a procedure similar to that described previously has been applied to analyze and predict the log P values of substituted pyridines and diazines, in which the fused N atom in the ring is dealt with as being a substituent (Yamagami et al., 1995). The procedure works well with certain approximations, but indicates that further elaboration is required for heteroaromatic compounds. The log P value of a number of zwitterionized di- to pentapeptides at the isoelectric point has been analyzed with the use of free-energy related parameters under defined conditions (Akamatsu and Fujita, 1992). The  parameter of the side chain of the amino acid units is that for aliphatic substituents (Iwasa et al., 1965; Leo et al., 1971). Along with the term for the steric effect of the side chain in terms of a variation of Es, correction terms for polar side chains interacting with the backbone –CONH– structure and the

Chapter | 58  Hydrophobicity as a Key Physicochemical Parameter of Environmental Toxicology of Pesticides

-turn formation are included (Akamatsu and Fujita, 1992). A similar set of aliphatic parameters is used in the analyses of the log P values of primary, secondary, and tertiary amines and quaternary ammonium ions pairing with picrate (Takayama et al., 1985).

58.3.5  Computer-Aided Procedures To make a more comprehensive procedure that covers a wide variety of compounds as aliphatic, alicyclic, (hetero)aromatic, and various combinations, it should be computerized with an access to a large database of reliably measured log P values. Various “rules” governing the structural contribution to log P, and correction terms for various intramolecular interaction features specific to each series should be incorporated into the program software, and calculated values should be immediately compared with the measured values in the database whenever available. The software could be constructed so as to recognize structural features of compounds when it is input into the computer and to output the calculated log P value “automatically” after data processing according to the rules and corrections covering any types of structural features. There are quite a few programs based on this type of concept (Sangster, 1997b). One of them, the “fragmental method,” was first proposed by Nys and Rekker (1973). They have been developing software for their own fragmental method (Rekker and Mannhold, 1992). With a procedure different from that used in the Rekker method, Leo and Hansch developed another fragmental method, the software of which is called CLOGP (Hansch and Leo, 1979, 1995; Leo, 1991, 1993; Leo et al., 1975). In this section, the CLOGP procedure is briefly described. In principle, the fragmental method is based on a “correlation” equation such as the following to elucidate the measured log P value:

log P 

∑ an fn  ∑ bm Fm .

(18)

Here, a is the number of occurrences of molecular fragment f of type n, and b is the number of occurrences of correction factor F of type m. The fragmental hydrophobicity index f differs from the  value for a certain substituent X. For instance, the log P of chlorobenzene is conceptually expressible by either of the following equations, although the f(C6H5) value is algo-rithmically dealt with as being divisible into smaller fragments in the CLOGP:

log P(PhCl)  log P (PhH)  (Cl/PhH ),

(19)



log P(PhCl)  f (C6 H 5 )  f (Cl).

(20)

Because log P(PhH) in Eq. 19 is expressible as the sum of f(C6H5) and f(H), the relationship between the  and f values can be represented as

(Cl/PhH )  f (Cl, aromatic)  f (H, aromatic). (21)

1237

Table 58.2 lists the f values of some representative substituents as well as the F values of typical correction features (Leo, 1998). As indicated, an “isolating carbon” (IC) atom is defined as the one not doubly or triply bonded to a heteroatom. An IC can be bonded to heteroatom inside an aromatic ring, and one IC can be multiply bonded to another. The IC and hydrogens attached to it (ICHs) are considered hydrophobic fragments. All atoms and groups of covalently bonded atoms, which are left after removing ICs and ICHs, are considered polar fragments. The polar fragments do not contain ICs, but each is connected to ICs with one or more bonds. The f value is assigned first to each fragment. Depending on the bond environment, such as aliphatic and aromatic as well as benzyl and vinyl, the f value of polar monovalent fragments is found to vary, with the aromatic value highest and the aliphatic value is lowest. Vinyl and benzyl values are intermediate. This difference is probably due to variations in the degree of delocalization of electron pairs of the polar fragment. The correction factors, F, for six types of fragment interactions are noted in the aliphatic systems. When halogens, X, and hydrogen-bonding fragments, Y, are located either geminally or vicinally in such arrangements as X– C–X, X–CC–X, X–C–Y, X–CC–Y, Y–C–Y, and Y–CC–Y, the F value for “proximity polar” interactions is assigned to each of them as positive corrections with detailed rules depending on their structural features. For aromatic systems, electronic interactions among substituents modifying their hydrogen-bonding capability, such as those described in the preceding section, are incorporated after modifications/simplifications. The “ordinary” electronic effect of a certain substituent designated by the 0(o, m, p) in Eqs. 16 and 17 is simplified/ approximated here as being expressible as a single  value, some of which are listed in Table 58.2. The  value of substituents/fragments, which is not easily accessible directly by the analyses according to such correlation equation as Eqs. 13–17, is estimated/calculated so that each of the / combinations (or YX products) makes up the difference between the measured value and the simple fragment sum. The  values of some fragments are also shown in Table 58.2. Provisions are also made for multiply substituted compounds, including heteroaromatic systems. Thus, the susceptibility  of each substituent to multiple interaction is not additive but attenuated starting with the greatest / combination of fragments (substituents) regardless of whether they are present as substituents or as fragments fused in a heteroaromatic ring. For ortho disubstitutions, besides the “ordinary”  interaction, a negative correction factor is assigned when the effect is regarded as a twisting of one of the substituents out of the ring plane, whereas a positive correction factor is defined for the internal hydrogen bond formation. In addition, the CLOGP program uses correction factors for the bond flexibility in aliphatic systems. For chain compounds, the total correction is made by a negative unit

Hayes’ Handbook of Pesticide Toxicology

1238

Table 58.2  Hydrophobic Fragment Constants (f) and Correction Factors (F) f(al.)a

Fragment H

0.23

F

f(ar.)b 0.23

c

d

Fragment

f(al.)a

f(ar.)b

c

d

—e

—e

–O–

1.82

0.61

0.17

0.50

f

0.38

0.37

0.28

0

–N5

2.37

1.12

0

0.61

Cl

0.06

0.94

0.28

0

–S–

0.79

0.03

0

0

Br

0.20

1.09

0.28

0

SO 2

3.13

2.17

0.70

0.45

I

0.59

1.35

0.28

0

–NH–

2.15

1.03

0

1.08

NO2

1.16

0.03

0.60

0

–CO–

1.84

1.09

0.51

0.27

OH

1.64

0.44

0

1.06

CO 2

1.45

0.56

0.51

0

SH

0.23

0.62

0

0.50

–CH5N–

1.20

1.03

0

0.61

NH2

1.54

1.00

0

1.08

–CONH–

2.71

1.81

0.32

0.72

NHCONH2

2.18

1.07

0

1.08

–OCONH–

1.79

1.46

0.17

0.50

SO2NH2

2.37

1.61

0.35

0.88

–NHCONH–

2.18

1.57

0

1.08

CF3

—e

1.11

0.49

0

–CON(A)g

3.14

2.80

0.51

0.27

CN

1.27

0.34

0.65

0

–OPO(OA)2g

2.29

1.71

0.17

0.80

COOH

1.07

0.03

0.32

0.35

–N5h

—e

1.14

0.90

0.30

CONH2

1.99

1.26

0.32

0.60

5C5i

—e

—e

Aliphatic features

F

Aliphatic features

F

Chain branch

0.13

X–C–X

0.6–2.8j k

Group branch

0.22

X–C–C–X

0.28

Double bondm

0.09

X–C–Y

0.9–2.7n

Triple bondm

0.50

X–C–C–Y

0.35–0.45o

Chain bondp

0.12

Y–C–Y

0.26–  0.42q

Ring bondr

0.09

Y–C–C–Y

0.15–  0.26q

0.20

0.13

Aromatic features

F

Internal H bonding

0.63

Ortho effect

0.28l

Electronic interaction among fragments

∑  s

a

Aliphatic. Aromatic. Electron-withdrawing effect of aromatic fragments on others regardless of their positions.

b c

∑

 of other aromatic fragments. Susceptibility to the Not applicable. f Trivalent nitrogen. g “A” means to be connected to the aliphatic IC. h Nitrogen fused in heteroaromatic rings such as pyridine. i Isolating carbon. j Increases with the number of geminal halogens (X) from 2 to 4. k Multiplied by the total number of halogens (on both sides of CC) minus unity. l At least one of the ortho pair is capable of H bonding. Multiplied by a number defined for each combination of ortho pairs. m For isolated bonds. n Varies depending on the type of Y (H-bonding fragment) and the number of geminal X. o 0.45 when X  F. p Multiplied by the number of bonds connecting fragments minus unity. d e

∑

f (Y), which is usually negative to give positive corrections; varies according to structural features of Y. Multiplied by Multiplied by the number of bonds. s See Table 58.4. q r

Chapter | 58  Hydrophobicity as a Key Physicochemical Parameter of Environmental Toxicology of Pesticides

F value multiplied by the bond number connecting (or outside of) fragments minus unity (not counting those to H). For alicyclic compounds, the bond flexibility correction is made by a less negative factor. The branching structures in the alkane chain at ICs and at polar fragments are also considered as negative correction factors. Table 58.3 shows the CLOGP components of methomyl, an oxime carbamate insecticide (Leo, 1998). The polar fragment, including the amide H, is enclosed by the broken line, and three CH3 groups, each including an IC and three ICHs, connected to the polar fragment, are circled in the structural formula of methomyl. The CLOGP estimation is in a good agreement with that measured by Drabek and Bachmann (1983). Table 58.4 is that for the more complicated triazine herbicide, terbutryn (Leo, 1998). The six polar fragments numbered from 1 to 6 are shaded circles, whereas three aromatic ring ICs and seven aliphatic ICs with respective ICHs (five CH3, one CH2, and a quaternary C) are just circled. In Table 58.4, the f value of secondary (2°) amino (1.030) and thio (0.030) fragments is that for the aromatic substituents. The f value of aromatic ring IC (0.130) is defined as being lower than that of aliphatic IC (0.195). In this molecule, the correction for the bond flexibility is made for each of the “side chains.” The NH–tertiary butyl substituent is considered double-branched, whereas the NH–ethyl is a straight chain. Electronic interactions are considered between the three substituents and the three–N fragments fused in the triazine ring as mentioned before. In terbutryn, the electronic interactions are assumed to be “faded with use” in terms of the YX product. For instance, each of the fused “aza” nitrogens and secondary amino groups (X) undergoes the electronic effect Y of other fragments according to the susceptibility X and its fading factor. The situation is illustrated in the lower half of Table 58.4. The ortho correction factor of 0.400 is empirically assigned because the measured log P value of 2-methylthiopyridine (Yamagami and Fujita, 1995) deviates positively from the CLOGP similar to this value. The value of 3.129 here for terbutryn is calculated using the newest version of the program (Leo, 1998) and differs from that in earlier publications, 3.73 (Hansch and Leo, 1995; Leo, 1991), but it is closer to the more recently measured value of 3.38 (Liu and Qian, 1995). With the preceding types of improvements for the estimation of the log P values of complicated compounds, the latest version of the CLOGP program seems to work very well. The correction factors are physicochemically as well as empirically reasonable enough. The elaboration in the treatment of the electronic interactions among aromatic substituents to assume the fading effect has been observed to also work well for the calculation of the log P of such candidate azole fungicides as shown in Table 58.5 with the five-membered heteroaromatic system (Kataoka et al., 1989).

1239

Table 58.3  CLOGP Example Calculation of Methomyl CH3 CH3

S

C

N

O

O

H

C

N

CH3

Fragment types, Components in CLOG correction factors

Estimated values

Polar fragment

Thioiminocarbamate 1.850 (f)

Isolating Cs

3  aliphatic ICs

0.585 (3  0.195f)

Ex-fragment Hs

9  H on ICs

2.043 (9  0.227f)

Ex-fragment bonds (3  1)  bond factor 0.240 (2  0.12F) Measured  0.60a a

∑ af  ∑ bF  0.538 (CLOGP)

Drabek and Bachmann (1983)

Although a number of correction factors are required, this is a matter of course if one can understand that there are various types of interactions between solutes and solvents as well as between substructures within single molecules that can affect the log P value. Further elaboration is needed because a number of factors have yet to be taken into account (Leo, 1998). It is recommended not to evaluate the computer-aided system in terms of its computational efficiency/simplicity but to judge it in terms of the physical organic background of the program.

58.4  Physicochemical significance of log P in environmental quantitative structure-activity relationships 58.4.1  Behavior in Soil The behavior of pesticides in soil is governed by adsorption, movement, vaporization, and degradation (Arnold and Briggs, 1990). The adsorption of organic chemicals to soil and sediment plays a very important role in their transport and mobility in the environment. The adsorption is an exothermic process, and a strong adsorption is observed in a low temperature range. The soil adsorption coefficient, Kd, in a soil/slurry-water system is expressed by the ratio of the amount of compounds adsorbed in soil (g/g soil) to the concentration in water (g/ml) at an equilibrium state under defined conditions. The soil adsorption of organic un-ionizable pesticides has been shown to be expressible by the hydrophobic/ hydrophilic balance parameterized by the partition between

Hayes’ Handbook of Pesticide Toxicology

1240

Table 58.4  CLOGP Example Calculation of Terbutryn

1 S

CH3

2 H3C

CH3 C

H3C

N

N H 3

N

N 4

6 H2 C N H 5

CH3

Fragment types, correction factors

Components in CLOGP

Estimated values

Polar fragments

2  2° amine

2.060 (2  1.030f)

Polar fragments

3  aromatic (fused) N

3.420(3  1.140f)

Polar fragment

sulfide, –S–

0.030 (f)

Isolating Cs

7  aliphatic ICs

1.365 (7  0.195f)

Isolating Cs

3  aromatic ICs

0.390(3  0.130f)

Ex-fragment Hs

17  H on ICs

3.859 (17  0.227f)

Ex-fragment bonds

7  bond factor

0.840 (7  0.12F)

Ex-fragment branch

1  chain branch

0.130(F)

Ex-fragment branch

1  group branch

0.220 (F)

Electronic interactions

∑  ,

Ortho correctiona

Between –S– and 5N–

3.755 (F)

see below

0.400 (F)

∑ af  ∑ bF  3.129 (CLOGP)

Measuredb  3.38

Composition of electronic interactions Composition of effects towards Xc

Fading factor of each  productc

Totalc

Substituent (X)

X

X

1

–S–

0.00

0.00

2

5N–

0.90

0.30

(4-N,6-N)  (2-N)

1.0, 0.3

0.351

3

–NH–

0.00

1.08

(2-N,4-N,6-N)  (3-NH)

1.0, 0.3, 0.09

1.351

4

5N–

0.90

0.30

(2-N,6-N)  (4-N)

1.0, 0.3

0.351

5

–NH–

0.00

1.08

(2-N,4-N,6-N)  (5-NH)

1.0, 0.3, 0.09

1.351

6

5N–

0.90

0.30

(4-N,6-N)  (6-N)

1.0, 0.3

0.351

0

∑ (Aromatic electronic interaction)  3.755 a

The positive correction may be due to a 1 :1 hydration in the octanol phase (Fujita, 1983). From Liu and Qian (1995). c After the  value of each of the “substituents” parenthesized is multiplied by the  value of X, each of the  products is weighted by each of the “fading factors” respectively in the descending order and the weighted  products are summed up, leading to “total.” b

Chapter | 58  Hydrophobicity as a Key Physicochemical Parameter of Environmental Toxicology of Pesticides

organic solvent and water or the chromatographic retention (Hance, 1967). Thus, for a certain soil sample, the log Kd value is observed to be almost linearly correlated with the log P value for a related series of compounds. For instance, Uchida and Kasai (1980), using a local soil sample in Japan, obtained the following linear relationship between the log Kd values of a rice blast fungicidal isoprothiolane (VI: R1  R2  i-Pr) and related compounds and their log P values:

Table 58.5  CLOGP Calculation of Triazole Fungicides (CH2)n H

n  12,

s  0.150,

N

Uchida et al. (1982b) also found that the log Kd values calculated from Eq. 22 for structurally unrelated buprofezin (VII), an insect growth regulator, and flutoranil (VIII), a rice sheath blight fungicide, agree well with experimentally estimated Kd values using the same soil sample as before. The basic mechanism of the soil adsorption of organic compounds from the water phase had been recognized earlier to be that involved in the distribution between soil organic matter and water (Goring, 1962). Thus, the “soil organic-carbon adsorption coefficient” or Koc should be normalized by the proportion of organic components in soil samples, Foc, as determined in a separate experiment and defined by R1OCO

S

N N

S

R2OCO

N

tert-Bu

S

VI

VII CF3 H N

O



N

Measureda

CLOGPb

2-Cl

4

2.68

2.63

2,5-Cl2

4

3.24

3.34

4-C1

5

3.03

3.19

Kataoka et al. (1989). Leo (1998).

b

sediments, collected from various districts in the United States, yielding log K oc  0.76 log P  1.66, n  19,

s  0.27,

r  0.93.

(24)

These relationships for a number of individual series of compounds were comprehensively examined by Sabljic et al. (1995). They derived the following equation for a number of organic compounds belonging to such classes as acetanilides, carbamates, phenylureas, phosphates, triazines, triazoles, and uracils, including practically used agrochemicals:

iso-Pr

log K oc  0.47 log P  1.09,

O



X

a



iso-Pr

O

N



(22)

r  0.980.

N

HO

X

log K d  0.41(0.06) log P  0.40(0.24),

1241

VIII

K oc  K d /Foc .

(23)

The definition assumes that pesticide adsorption by soils is entirely due to organic matter, even though the organic matter is a complex mixture of carbon, hydrogen, and nitrogen compounds, which acts as a nonpolar film coating soil particles. The Koc value is relatively constant for a particular compound among soil samples from different origins. Briggs (1981b) reported that the Koc value of herbicidal phenylureas measured with English soils agrees well with that measured with Australian soils. Dzombak and Luthy (1984) showed that the Koc value of chlorobenzenes hardly depends on soil samples having various fractional organic matter ranging from 0.15 to 33%. Oliver (1987) determined Koc values for chlorinated hydrocarbons, including polychlorinated biphenyls with suspended

n  216,

s  0.425,

r  0.826.

(25)

In Eq. 25, log Koc varies from 0 to 5 and log P from 1 to 8. They proposed that this equation can apply to the prediction of log Koc values of nonmeasured agrochemicals irrespective of origin of the soil samples. In fact, Eq. 25 is very similar to the following equation, which was derived for phenylurea-type herbicides independently (Liu and Qian, 1995) using soil samples from China: log K oc  0.59(0.15) log P  1.18(0.43),

n  9,

s  0.152,

r  0.961.

(26)

The mobility of pesticides in soil is important in governing their persistency as well as downward movement to pollute the groundwater and lateral movement to pollute surface water. Mobility can be estimated using soil column chromatography or soil thin-layer chromatography. Uchida

Hayes’ Handbook of Pesticide Toxicology

1242

and Kasai (1980) studied the mobility of isoprothiolane and its analogs (VI) using a soil column chromatography and expressed it as log , where  is the ratio of the volume of soil packed in the column to the volume of the aqueous phase required to elute the solute i.e., [  Vsoil/Veluent They derived the following equation: log   1.44(0.17) log K d  0.45(0.33),

n  11,

s  0.193

r  0.986.

(27)

Equation 27, together with Eq. 22, indicates that the greater the log P, the lower the mobility of the compounds in soil. Briggs (1973) obtained similar results for herbicidal phenylureas with soil thin-layer chromatography. Their mobility in terms of the RM (Boyce and Milborrow, 1965) normalized by the content of organic matter was related to the log Koc value irrespective of the source of soils. Helling (1971) measured the mobility of ionic/ionizable herbicides such as dicamba, 2,4-dichlorophenoxyacetic acid (2,4-D), fenac, picloram, and diquat by thin-layer chromatography with 14 kinds of soil samples. The mobility of ionic/ionizable compounds was not simply related to the log Koc value. Their soil adsorption mechanism is not a simple hydrophobic partitioning into the soil organic matter from the aqueous phase, but includes various interactions. According to Wauchope et al. (1992), these interactions include (a) binding of cations to negatively charged sites on clay surfaces (a very strong interaction), (b) binding of anions to soil anion-exchange sites (a very weak interaction), and (c) specific chemical binding mechanisms such as the phosphate-fixation-like binding of glyphosate and the arsenicals to soil metal oxides. In many cases, anionic and cationic pesticides, which give very low and very high Kd values, respectively, have no reported soil adsorption values, probably because the extreme values involved are difficult to measure. The vapor losses of volatile pesticides from soil to air depend primarily on the air/soil distribution constant, KAS. Vaporization from water is conveniently estimated by the Henry’s law constant, KAW (air/water). In the soil/water/ air system, the soil adsorption lowers the concentration of pesticides in the water phase and, in turn, in air. The air/ wet soil distribution could be approximated by the ratio of KAW and a soil/water distribution constant such as Koc The KAS value is thus a function of water solubility, vapor pressure, and soil adsorption. According to Arnold and Briggs (1990), the KAS constant is roughly expressible as a function of boiling point and log P value under denned conditions for such factors as the air/soil ratio, distribution of the pesticide in the soil sample, and climate. To a first approximation, pesticides in soil exhibit an exponential degradation according to the first-order kinetics. The degradation half-life, dT50, can be estimated from the reciprocal of the first-order rate constant. Degradation in soil occurrs, however, as a combination of mechanistically

complex processes. It generally includes abiotic and biotic processes. The abiotic degradation is due to chemical reactions such as hydrolysis, photolysis, air oxidation, and others. Depending on the structural feature, pesticides could suffer from various types of reactions. No common parameter such as log P alone is capable of describing a variety of reactivities. The reaction rate can be evaluated if experimentally established model systems are available. For biotic degradation of organic compounds, there have been quite a few efforts to establish the QSAR model correlation equations. With the use of acclimated mixed microbial cultures, Babeu and Vaishnav (1987) measured the 5-day BOD (biological oxygen demand in mmol/mmol chemical) of a wide variety of organic compounds, including alcohols, acids, esters, ketones, and aromatics. They examined the correlation of log(BOD) with various physicochemical parameters for 45 compounds. Their analysis, showing that the log(BOD) values fit well a correlation equation with quadratic terms of log(theoretical BOD), seems to be reasonable (n  45, r  0.862) among others. There would be an optimum theoretical BOD value for compounds to be most biodegradable. Molecules in which the number of carbon atoms is high necessarily have a high theoretical BOD value. They are often highly hydrophobic. The highly hydrophobic compounds, being trapped by cell membrane lipids, would not be easily incorporated into microbial cells. In fact, using the BOD data of Babeu and Vaishnav, Zakarya et al. (1993) formulated a correlation equation showing that the experimental BOD values are parabolically related to log P and linearly related to molecular volume (n  43, s  0.575, r  0.906). In contrast to the preceding studies, Dearden and Nicholson (1986) discussed the significance of the electronic structure of the molecule. The 5-day BOD value for various types of compounds, including amines, phenols, aldehydes, acids, halogenated hydrocarbons, and amino acids, supplied by the U.S. Environmental Protection Agency, Duluth, Minnesota, was shown to be highly dependent on a new electronic parameter that is expressible as the difference in the modulus of atomic charge across a key bond in the molecule, which could be attacked by microbials (n  79, s  3.459, r  0.993). In spite of the fact that the biodegradability is estimated under simplified conditions without soil, the QSAR model building needs much improvement. In reality, assignment of a single dT50 value to each pesticide is impossible under field conditions. Although it could be related to dT50 values from model systems, it is also highly sensitive to the type of soil with varying mineralogy, carbon and water content, and pH, and the distribution and activity of soil microbials as well as climate. Thus, it is almost impossible to build a comprehensive single QSAR model for the degradation of a variety of pesticide classes. Careful studies should be made under defined conditions perhaps on a series-toseries basis.

Chapter | 58  Hydrophobicity as a Key Physicochemical Parameter of Environmental Toxicology of Pesticides

58.4.2  Bioaccumulation Processes involved in the accumulation of environmental chemicals in various organisms through aquatic phases are generally classified into two types: bioconcentration and biomagnification (Connell, 1988). Bioconcentration is the accumulation of chemicals dissolved in water in fish and aquatic organisms through the gills and body surface directly. The bioconcentration factor (BCF) is defined as the ratio of the concentration of a chemical in an aquatic organism to that in the aqueous phase under steady-state conditions. The measurement of the BCF has been made with the average concentration of the chemical in the whole body absorbed through the gills, skin, and digestive tract of fish of small to moderate size reared in a sublethal aqueous concentration. Sometimes, the BCF is estimated as that for the lipid content of fishes. Following the work of Neely et al. (1974) showing that the log value of the BCF of nonpolar compounds can be correlated with their log P value, a number of examples have been accumulated. Mackay (1982) critically reviewed the BCF values and proposed that the fundamental process observed in bioconcentration is such that P and BCF are proportional; that is, the slope of the log BCF versus log P correlation should “theoretically” be close to unity. In the earlier publications (Neely et al., 1974; Veith et al., 1979), the slope was often lower than unity, even when dealing with nonpolar compounds. Mackay suggested that the lower slope was the result of overestimation (by calculation) of the log P of compounds of high molecular weight. After omitting compounds, the log P value of which are above 6 or are unreliable, compounds that are ionizable, and compounds that can act as surfactants, Mackay (1982) proposed the following equation for 43 compounds. These are mostly nonpolar such as lindane, DDT, and (polyhalogenated) aromatic hydrocarbons with log P values ranging from unity to 6.



log BCF  log P  1.32(0.25), n  43, r  0.974.

(28)

Equation 28 is valid as far as inert compounds having log P values lower than 6 are concerned. It also indicates that variations in fish species, with which BCF is measured, are insignificant in the general relationship between log P and log BCF values. The lipid content of the fish used in developing Eq. 28 has been estimated as being 5% (Connell and Hawker, 1988) and does not vary significantly among fish species. There are quite a few examples conforming to Eq. 28. Oliver and Niimi (1983) determined the BCF of chlorobenzenes in rainbow traut (Salmo gairdneri) as



log BCF  1.022(0.057) log P  0.632, n  11, r  0.993.

(29)

1243

In this and other correlation equations (Connell, 1988; Davies and Dobbs, 1984; Isnard and Lambert, 1988; Oliver and Niimi, 1985; Opperhuizen et al., 1985), the slope of the log P term is indeed close to unity and the intercept ranges from 0.5 to 1.3. Considerable deviations from the linear relationship represented by Eqs. 28 and 29 have been observed, however, for highly hydrophobic compounds with a log P value 6 (Bruggeman et al., 1984; Opperhuizen et al., 1985). Reduced membrane permeation (Opperhuizen et al., 1985), lowered lipid solubility (Banerjee and Baughman, 1991), and other possible reasonings have been proposed. To simulate this situation better, Bintein and Devillers (1993) proposed the following equation using the bilinear model developed by Kubinyi (1977) for a number of compounds, roughly one-third of which belongs to pesticides: log BCF  0.910 log P

(30) 7



 1.975 log(6.8  10 P  1)  0.784, n  154, r  0.950, s  0.347, F  463.5.

In Eq. 30, F is the ratio of regression and residual variances. The compounds were selected so that they are mostly inert and cover a wide range of log P values, ranging from unity to 9. In the second term on the righthand side, conventionally written as b[log(P  1)], the  value is supposed to correspond to the volume ratio between lipid and aqueous phases involved in the entire system for the manifestation of biological “activity.” For small P values, (P  1) is close to unity, so log(P  1) is 0, and Eq. 30 takes the form of Eqs. 28 and 29. For large P values, (P  1) is almost equal to P, so that log(P  1) is linear with log P. The value of – log  nearly corresponds to the log P value where the log BCF is maximum. Biphasic functions with linear ascending and descending sides and a rounded apical part are represented by this model. In Eq. 30, the positive slope for the ascending side is 0.910, and the negative descending slope is (0.910  1.975  1.065). Among the compounds included in Eq. 30, 24 compounds have a log P higher than 6; that is, they are covering a part of the apical region and the descending phase. In measuring the BCF values of compounds having a large log P value, one needs to carefully set the test period. Oliver and Niimi (1983) showed that 120 days are needed to attain the equilibrated steady state for polychlorobenzenes in fish. Hawker and Connell (1985a, b) suggested that half a year may be required to obtain the steady state for compounds with a log P of around 6 and about 10 years for those with a log P of about 8, and formulated a QSAR similar to Eqs. 28 and 29 for the BCF value under nonequilibrium conditions with a certain exposure period.

Hayes’ Handbook of Pesticide Toxicology

1244

Devillers et al. (1996) compared in detail the versatility of the bilinear model expressed by Eq. 30 with that of linear, quadratic, and polynominal correlation models published from different organizations. They selected 342 log BCF values for 181 compounds, which are mostly inert. Some values were independently measured in duplicate for a single compound. The selection criteria for these log BCF values required that the BCF data are obtained only after a steady state was established and that, if one or more values appeared out of line in a publication, all the data contained in that publication are not used. They concluded that the bilinear model represented by Eq. 30 is among the best in predictive performance in terms of the root mean square (rms) value for residuals between log BCF values experimentally measured and calculated from model equations. Banerjee and Baughman (1991) proposed another model for the nonlinearity of log BCF versus log P. They tried to rationalize the breakdown of the linear relationship not only for the highly hydrophobic compounds, but also for many azo dyes (multifunctionalized azobenzenes), the log P value of which is below 6 (mostly between 3.5 and 4.5) (Anliker and Moser, 1987). Their model considered the fact that large compounds of low lipid solubility such as azo dyes and polyhaloaromatic hydrocarbons, have lower than expected BCF values because of the difficulty in cavity formation in lipids. Thus, with an approximation in which the lipid solubility of compounds (generally unavailable) is replaced with the solubility in octanol, Soct (M), they proposed the following equation for a set of compounds. These include inert pesticides, aromatic and aliphatic (halogeno)hydrocarbons, and polar but mostly nonionized azo dyes, the log P values of which ranged from 1.5 to 8.3.



log BCF  1.13  1.02 log P  0.84 log Soct  0.0004(mp  25), n  36, r  0.95. (31)

In Eq. 31, mp is the melting point (in °C), which was intended to allow octanol solubilities for both liquids and solids to be included in a single equation (Valvani and Yalkowsky, 1980). For liquids, mp is regarded as 25 to remove the entire term. For small compounds, the log Soct and (mp – 25) terms in Eq. 31 tend to be constant and Eq. 31 takes the same form as Eq. 29. Most of the compounds included in Eq. 31 are solids. Without the log Soct and (mp – 25) terms, the correlation was much poorer (r  0.73). In the region of log P 6, even though the log P increases, the BCF value could decrease because of a decrease in the lipid (octanol) solubility, resulting in the descending phase in the log BCF/ log P relationship. The previous relationships seem to be in accord with the low fish toxicity of an insecticide of nonester-type pyrethroids, silafluofen (IX) (Sieburth et al., 1990). The

log P value of this compound has been estimated as being about 10 (Okimoto et al., 1994). Its uptake in fish should well be very low. Me C2H5O

Si

F

(CH2)3

Me O IX

For less inert compounds, including pesticides in which certain reactive/vulnerable functions are required for their biological activity, the slope has been observed to be significantly lower than unity. Moreover, the correlation is often of a lower statistical quality (de Wolf et al., 1992; Niimi et al., 1989; Oliver and Niimi, 1985; Opperhuizen and Voors, 1987). Such deviations from Mackay’s postulate (Mackay, 1982) have been attributed to relatively high rates of biotransformation (de Bruijn and Hermens, 1991; Opperhuizen and Voors, 1987; Southworth et al., 1980). Uchida et al. (1982a) measured the BCF value of isoprothiolane and its analogs (VI) with the killifish (Orizias latipes) and proposed the following: log BCF  0.65(0.17) log P  1.17(0.62), (32) n  9, s  0.197, r  0.962, F  85.5. The reason that the coefficient of the log P term is smaller than unity is attributable to the fact that the test compounds have two hydrolyzable ester groups. Uchida et al. also measured the rate of disappearance of the compounds in the entire system. The rate followed approximately the zero-order kinetics, and the rate constant, log k, tended to increase with decreasing bulkiness of the ester substituents in terms of the STERIMOL width (Verloop, 1983) and with increasing log P value. This was taken to indicate that the hydrolytic degradation would occur under conditions “reacting” with fish. The “apparent” log BCF term in Eq. 32 should be compensated for the degradation effect corresponding to the log k value so that the size of the log P term under the “real” conditions would be closer to unity. de Bruijn et al. (1993) examined the effect of biotransformation of a set of organophosphorus insecticides on bioconcentration. The compounds belonged to O, O-dimethyl-O-phenyl phosphorothioates in which various substituents such as CN, NO2, SMe, and halogens are located at the 2-, 4-, and 5-positions of the benzene ring either singly or multiply. Their log BCF values were measured using guppies, from which the following equation was derived:



log BCF  0.80(0.12) log P  0.45, n  12, s  0.35, r  0.910.

(33)

Chapter | 58  Hydrophobicity as a Key Physicochemical Parameter of Environmental Toxicology of Pesticides

One of the most important metabolic pathways of dimethyl phosphorothioates has been shown to be the demethylation of one of the two methyl groups by glutathione (Fukami and Shishido, 1966). Thus, the rate of demethylation under pseudo-first-order conditions, k (min1 mg protein1), was measured using a glutathione S-methyltransferase preparation. Introduction of the log k term into Eq. 33 yielded the following:



log BCF  0.94(0.08) log P  0.63(0.14) log k  2.31, (34) n  12, s  0.21, r  0.971.

The improvement in the correlation quality is significant and the slope of the log P term is close to unity in Eq. 34. The preceding two examples indicate that, if an appropriate term for the vulnerability is incorporated, then Mackay’s postulate is valid even for the unstable series of pesticides. Mackay’s postulate has also been observed for bioconcentration in mollusks, daphnias, and aquatic microbes (Baughman and Paris, 1981; Geyer et al., 1982; Hawker and Connell, 1986). The structure–activity relationships for biomagnification are not as well established as those for bioconcentration. A number of mechanisms, which are not well understood, are involved in the entire process of biomagnification, which occurs through the food chain. However, it is highly probable that the biomagnification is also able to be significantly related to the octanol/water partition coefficient (Esser, 1986). Davies and Dobbs (1984) indicated that the uptake of chemicals from both food and water results in tissue concentrations comparable to those resulting from water alone. It should be mentioned that the uptake of pesticides from food is far less important than the uptake from water and that only a part of the residue present in the lower level biota is transferred to the higher level of the food chain (Ellgehausen et al., 1980).

58.4.3  Aquatic Toxicity The aquatic toxicity of “simple” organic compounds has been recognized as being closely related to their lipophilicity (hydrophobicity). Overton and Meyer independently proposed the “lipoid theory of narcosis” about 100 years ago (Lipnick, 1989), narcosis being considered as a toxic effect that can be lethal. The modern formulation of the aquatic toxicity in terms of the QSAR was, however, first published by Hansch and Dunn (1972). They found that the earlier toxicity data of various sets of homologous alcohols or miscellaneous inert compounds, in terms of the minimum narcotic and the minimum lethal concentrations, can mostly be explained by a single parameter log P with a slope close to unity. Subsequently, a number of hypotheses of aquatic toxicology were combined with the QSAR concept. Thus, Könemann (1981) investigated the QSAR of “environmental pollutants” of structurally heterogeneous

1245

organic compounds such as methyl- and chlorobenzenes, aliphatic chlorohydrocarbons, and alcohols, among others. For the 14-day LC50 (M) value in the guppy (Poecilia reticulata), the following equation was derived:



pLC50  0.871 log P  1.13, n  50, r  0.988, s  0.237.

(35)

In this set of compounds, the pLC50 (M) value varies from 6.15 for pentachlorobenzene (log P  5.69) to 0.24 for diethyleneglycol (log P  1.30). The selection of compounds included in Eq. 35 was intended so that they are rather stable and unreactive (inert) as well as un-ionizable. Thus, pesticides that are “reactive” to “specific” targets in vivo are not included. With a larger (4 day) toxicity data set against juvenile fathead minnows (Pimephales promelus) from the U.S. Environmental Protection Agency, Duluth, Minnesota, the following equation was derived (McCarty et al., 1992):



pLC50 (M)  0.90(0.04) log P  1.29(0.12), (36) n  150, r  0.959.

The compounds included in Eq. 36 are inert and cover the range of log P similar to that in Eq. 35. They are rather stable, belonging to halogenated aliphatic hydrocarbons, ethers, (halogen-substituted) alcohols, esters, ketones, phthalates, substituted benzenes, and several pesticides (mostly rather stable herbicides). For inert organic compounds, the log P value of which is very high (5–6), the pLC50 value has been observed to be lower than expected by such linear correlations as Eqs. 35 and 36, probably because of the limit in the water solubility (solubility cutoff) to be partitioned into the fish body required for the symptom of toxicity. This situation was confirmed by Veith et al. (1983), who formulated a bilinear correlation of excellent quality such as the following for the 4-day toxicity to the fat head minnow of five classes of unreactive compounds similar to those included in Eq. 36:



(37) pLC50 (M)  0.94 log P 5  0.94 log(6.8  10 P  1)  1.25, n  65, r  0.999.

Although the detailed mechanism of the acute fish toxicity is not completely clear, the observed symptoms caused by inert and unreactive compounds strongly suggest a mechanism categorized as general narcosis (Lipnick, 1990). In Eqs. 35–37, the coefficient of the log P term is close to unity and the intercepts are almost equivalent to one another irrespective of the test organism and the range of test compounds. A number of similar correlations have been accumulated for sets of inert compounds (Cronin and Dearden, 1995; Ikemoto et al., 1992; Lipnick, 1990). Thus, the general narcotic (analgesic) potency is dependent only on the

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overall hydrophobicity and not on the specificity of the chemical structure. Because every compound is supposed to exert at least this type of narcotic activity, the QSAR correlations similar to Eqs. 35–37 are considered to predict the minimum toxicity or the “baseline toxicity” (Lipnick, 1990) of any organic compound to aquatic biota, unless the compound is biodegradable and therefore less toxic. The toxic effects of mixtures of compounds from a single group in terms of the mechanism (type) of toxicity should be concentration additive. Concentration addition means that the LC50 of a mixture is observed at the sum of concentrations, c, of individual compounds (as the fraction of their own LC50) being 1.0 (∑ c/LC50  1.0) . The compounds included in Eq. 35 have been shown to be concentration additive, that is, to exert their effect by an equivalent mechanism (Könemann, 1980). Veith and Broderius (1987) noted that a number of compounds that appear to produce narcosis are significantly more toxic than the baseline toxicity. They are more polar and often have weakly acidic and/or hydrogen-bond donor groups, such as substituted phenols, mostly existing as the nonionized form under experimental conditions, and substituted anilines. This structurally heterogeneous set of compounds has been examined by the concentration additivity test, showing that these phenols and anilines indeed belong to a group exerting a narcotic syndrome [Type II (or polar) narcotic syndrome] differing from that of the unreactive inert compounds [Type I (or non-polar) narcotic syndrome]. For phenols and anilines substituted both singly and multiply, by alkyl, alkoxy, halogen, NO2, the following equation was formulated with the 4-day LC50 value against juvenile fathead minnows:



pLC50 (M)  0.65(0.07) log P  2.29(0.22), n  39, r  0.95. (38)

Prior to this, Saarikoski and Viluksela (1982) analyzed the 4-day LC50 value of variously substituted phenols against guppies (P. reticulata) at pH 6, 7, and 8. The substitution pattern of phenols was carefully selected so that the collinearity between the log P and  log Ka values is as low as possible.  log Ka, the difference between the log Ka and the reference log Ka of unsubstituted phenol, was used as a parameter for the electron-withdrawing effect of substituents. They also corrected for the effect of ionization on the LC50 value in terms of an “effective” concentration of the neutral form defined empirically (Saarikoski and Viluksela, 1981). For phenols substituted with Me, Cl, t-Bu, OMe, OH, and NO2 groups singly or multiply, Saarikoski and Viluksela derived the following equations:



pLC50 (M; corrected for pH 6)  0.67 log P  0.19 log Ka  2.33, n  19, r  0.985, s  0.17.

(39)



pLC50 (M; corrected for pH 7)  0.71 log P  0.19 log K a  2.23, n  19, r  0.978, s  0.21.

(40)

Because Eqs. 39 and 40 are practically identical, they can serve to predict the toxicity at any pH from 6 to 8. It is interesting to note that, for phenols, the  log Ka value of which is so low that the effects of ionization on the LC50 and electron withdrawal of substituents are not great, Eqs. 39 and 40 are very similar to Eq. 38. For phenols with more acidic hydrogen than anilines, the electron-withdrawing effect of substituents on the OH, which may hydrogenbond with possible taget site(s), seems to be accounted for explicitely in Eqs. 39 and 40. Equations 38–40 probably reflect the feature of Type II (or polar) narcotics, in which the slope of the log P term is lower and the intercept is higher than those of the nonpolar narcotics. The higher intercept may result from the hydrogen-bond donor group in phenols and anilines enhancing the interaction with the target site(s). The lower slope could mean that there is an increasing shift toward nonpolar narcosis with increasing hydrophobicity. For nonpolar narcosis, the slope is higher. However, the toxity of nonpolar narcotics is lower than that of polar narcotics in the region of low log P values. Thus, the shift of the narcosis type could make the slope higher. A possible reason for this may be related to differences in distribution so that compounds with a high log P value will show an increasing tendency to accumulate in the lipid phase (Cronin and Dearden, 1995). In Eq. 38, such phenols as Cl5- and 2,4-(NO2)2-derivatives are not included because these compounds do not share the concentration additivity with those included. In Eqs. 39 and 40, phenols with 2,5- and 2,4-(NO2)2-substitution patterns are not included because of their much higher toxicity than that predicted. The structural characteristics of these outliers are shared by the pesticidal phenols mostly acting as uncouplers with the mitochondrial oxidative phosphorylation. They are neither Type I nor Type II narcotics, but exert a toxicity about 10-fold higher than that predicted by the Type II correlation equations. A number of compounds with biologically and/or chemically “reactive” sites have toxicities considerably greater than those predicted for either nonpolar or polar narcosis. Various pesticides with specific structural features and/or specific functional moieties to inteact with respective target site(s) are such compounds not categorized into narcotics. These compounds are often defined to show “reactive toxicity” (Hermens, 1990). Ikemoto et al. (1992) examined the excess toxicity exerted by various pesticides against killifish (Oryzias lapipes) over the baseline toxicity which was formulated for nonpolar inert compounds such as homologous alkanols, chlorobenzenes, and alkylbenzenes. Photosynthesis-disrupting herbicides, such as diuron, and chitin-synthesis-disrupting larvicides, such as diflubenzuron (IV: X1  X2  F,

Chapter | 58  Hydrophobicity as a Key Physicochemical Parameter of Environmental Toxicology of Pesticides

Y  Cl) and buprofezin (VII) were almost on the line for the nonpolar narcotic compounds. Their targets do not exist in fish. Such neurotoxic insecticides as DDT, dieldrin, fenvalerate, and lindane are more toxic than the baseline by 1.2–2.5 log units, and the respiration-inhibiting rotenone is higher than 3 log units more toxic. The neurotoxicity and respiratory toxicity are probably common between fish and insects. Besides the specific biochemical mechanisms exerted by pesticides, a variety of chemical reactivity mechanisms such as electro- and nucleophilic, redox, and free-radical processes are thought to be involved in the interactions of various types of toxicants with biological systems. Especially important are toxicants expected to work as electrophiles which can react with nucleophilic groups such as NH2, OH, and SH in deoxyribonucleic acid (DNA) and proteins. Hermens (1989) classified the reaction mechanisms of nucleophilic groups in biological systems with electrophilic toxicants into (a) nucleophilic displacement reaction, (b) addition to carbon– oxygen double bonds, and (c) addition to activated carbon– carbon double bonds (the Michael-type addition). He also surveyed various molecular substructures present in possible toxicants where these types of reactions might be responsible for their unwanted activity (Hermens, 1990). In this type of reactive toxicity, one should not expect simple relationships between toxicity and hydrophobicity even when accompanied by electronic parameters such as those represented by Eqs. 35–40. For instance, Hermens et al. (1985) derived the following equation for the 14-day toxicity to guppies of 15 alkyl, alkenyl, acylmethyl, and benzyl halides:



(41) pLC50 (M) 1  0.224 log P  1.32 log(2484  k )  10.05, n  15, r  0.956, s  0.39.

The pLC50 value was only very poorly correlated with the log P value alone (r  0.41), but the correlation was much improved when the pseudo-first-order rate constant, k (in min1), with a model nucleophile (4-nitrobenzylpyridine) was included. The second term on the right-hand side of Eq. 41 means that the variation in the toxicity is biphasic with respect to log k. For compounds with very low reactivity, that is, for those with very small k or very large kl values, the value of 2484 within the parenthess can be neglected relative to the kl value. Therefore, the value of the negatively signed second term, being approximated by 1.32 log k, initially increases nearly linearly with increasing log k, but, in the region above log k 3.4 (k 1/2484), follows a plateau-like pattern. Examples of the use of experimentally estimated reactivity indices have been reviewed by Hermens (1990), who also analyzed the 14-day LC50 values of epoxides and aldehydes and formulated correlations somewhat similar to Eqs. 38 and 41, respectively.

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In the reactive toxicants, the term “reactive” encompasses a wide spectrum of chemical processes as mentioned previously. Further classification must be made for hard (charge controlled) and soft (orbital controlled) interactions (Comporti, 1989) because relevant molecular descriptors are different between these interactions. Soft electrophiles cover one of the largest groups of anthropogenic toxicants exceeded only by narcotics. Soft electrophilic interactions of toxicants with biomolecules are regulated by the ability of toxicants to accept electron density represented by the “nucleophilic” or electron-acceptor superde-localizability, SN (the term “nucleophilic” is from the side of “biological” reactants), through the orbital interactions. The SN value on the activated unsaturated carbon atoms and the frontier charge of the lowest unoccupied molecular orbital, f(LUMO), along with the log P value, are important in predicting the toxicity of such soft (pro)electrophiles (Mekenyan and Veith, 1993). The “proelectrophilic” mechanism of the toxicity refers to compounds that are not direct-acting electrophiles but are metabolized to electrophiles such as primary and secondary allyl and propargyl alcohols (Lipnick et al., 1985). Veith and Mekenyan (1993) examined to extend this approach to a large set of -electron systems such as aromatic compounds, including variously substituted hydrocarbons, phenols and anilines, and unsaturated aliphatic compounds, including alkenols and alkynols. These were selected to represent a wide variation in  bonding and a variety of modes of toxic actions. The compounds include polar and nonpolar narcotics, uncouplers with oxidative phophorylation, and “reactive” electrophiles and proelectrophiles. Veith and Mekenyan derived the following equation for the 4-day LC50 value of 114 compounds against fathead minnow: pLC50 (M)  1.49(0.53)  0.56(0.04) log P (42)  13.7(1.7)S N (av.), n  114, r  0.90, s  0.43, F  238.7. SN(av.) is the superdelocalizability averaged for the values assigned to atoms included in the conjugated -electron system. It is expected to represent the global electronacceptor character of the molecule. The log P and SN(av.) parameters are orthogonal for these compounds. Equation 42 accurately predicts the toxicity of proelectrophiles, when the SN(av.) value is calculated for their metabolites. The authors (Mekenyan and Veith, 1993) further showed that the soft electrophiles included in Eq. 42 can be clustered according to their SN values by defining “isoelectrophilic windows” along the toxicity response plane. Nonpolar narcotics are located in the lowest SN region where toxicity varies almost only with hydrophobicity. Polar narcotics are more toxic than nonpolar narcotics at similar values of log P and the toxicity increase can be illustrated by higher SN

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values (by stronger electronic interactions with cellular soft nucleophiles). Highly reactive soft electrophiles, which have dissociable protons, act as uncouplers. Electrophiles without dissociable protons indicate symptoms of reactive toxicity consistent with covalent bonding. Perhaps because of their high specificity, the aquatic toxicology of pesticides has not been studied intensively in terms of the QSAR. There have been studies for sets of miscellaneous pesticides, but their toxicity is usually dealt with as that of QSAR outliers with higher potency than narcotic toxicants (Ikemoto et al., 1992). Exceptions are the organophosphorus insecticides (Hermens et al., 1987; Schüürmann, 1992) where a series of O,O-dimethyl phenylphosphorothioates are substituted at the 2-, 4-, and 5-positions on the benzene ring by various groups such as H, CN, NO2, Me, halogens, and SMe. The 14-day LC50 toward guppies were measured and analyzed for 12 analogs to give the following (Verhaar et al., 1994): pLC50 (M)  0.38(0.12) log P  27.9(7.78)p(PO) (43)  11.0(5.43), n  12, r  0.861, s  0.371, F  12.9. In Eq. 43, p(PO) is the bond order between the central phosphorus atom and the phenoxy oxygen calculated by MOPAC 6.0 (Stewart, 1990). This descriptor was believed to contain information about the strength of the corresponding bond reflecting the nature of the phenoxide as the leaving group and thus to have relevance for the phosphorylation step of acetyl-cholinesterase inhibition. The negative sign of this term suggests that the lower the PO bond order, the more easily the phosphorylation of the serine OH of acetylcholinesterase occurs, leading to the higher toxicity. The preceding study originally analyzed the LC50 data with the use of experimentally estimated reactivity indices such as the rate constant k with the model nucleophile, 4nitrobenzylpyridine (Hermens et al., 1987). The following equation was fotmulated for 10 analogs (two compounds were added later to formulate Eq. 43):



pLC50 (M)  0.34(0.10)∑  0.76(0.19) log k  3.57, n  10, r  0.912, s  0.29.

(44)

With the same set of compounds, the use of ∑  instead of log k yielded a poorer correlation. The significant contribution to toxicity by the demethylation rate, k, can be taken as a possible (but not definite) hint that the in vivo demethylation may be involved in the acute fish toxicity of this series of compounds (Schüürmann, 1992). There are an increasing number of aquatic toxicity QSAR publications other than those described previously in which quantum-chemical descriptors are used to illustrate

the electronic mechanism involved (Cronin et al., 1995; Schultz et al., 1995). The examples described in this chapter are those in which the QSAR procedure is successful in building models to elucidate the acute aquatic toxicity of the rather limited number of compounds included in the individual sets. The correlation equations could predict the toxicity of nontested compounds with related structures. However, there are a vast number of miscellaneous compounds the environmental behavior of which should be evaluated. Not only because of time constraints, but also because of limited financial resources, it is impossible to test all of the existing chemicals experimentally. For commodity chemicals, pharmaceuticals, and pesticides, it is better to predict adverse (environmental) toxicological behaviors before synthesis. In many of the preceding examples, the mode (or type) of aquatic toxicological action was more or less preestablished or categorized without much difficulty. For a large number of compounds, ranging from (non)polar narcotics, to reactive toxicants, and to those acting with specific biochemical mechanism, the ab initio identification of the mode of action becomes very difficult. Thus, attempts have been made to “identify” the mode of toxic activity according to substructures that are expected to participate in certain specific types of reactivity (Karabunarliev et al., 1996a, b; Mekenyan and Veith, 1994; Verhaar et al., 1992). The procedure is computerized with incorporation of the substructure search algorithm into expert systems. For sets of compounds searched for from the large data set, quantum-chemical descriptors are used to analyze the toxicity data quantitatively along with such log P values estimated with use of the CLOGP method (Leo, 1993).

Acknowledgments The authors would like to dedicate this chapter to Professor Corwin Hansch of Pomona College, Claremont, California, on his 80th birthday. They would like to express their sincere thanks to Dr. Albert Leo of Pomona College for his invaluable discussions as well as for his careful review of the manuscript.

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Chapter | 58  Hydrophobicity as a Key Physicochemical Parameter of Environmental Toxicology of Pesticides

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Schüürmann, G. (1992). Ecotoxicology and structure–activity studies of organophosphorus compounds. In “Rational Approaches to Structure, Activity, and Ecotoxicology of Agrochemicals” (W. Draber and T. Fujita, eds.), pp. 485–541. CRC Press, Boca Raton, FL. Shimizu, R., Iwamura, H., and Fujita, T. (1988). Quantitative structure–activity relationships of photosystem II inhibitory anilides and triazines: Topological aspects of their binding to the active site. J. Agric. Food Chem. 36, 1276–1283. Sieburth, S. M., Manly, C. J., and Gammon, D. W. (1990). Organosilane insecticides. I: Biological and physical effects of isosteric replacement of silicon for carbon in etofenprox and MTI-800. Pestic. Sci. 28, 289–307. Smith, R. N., Hansch, C., and Ames, M. M. (1975). Selection of a reference partitioning system for drug design work. J. Pharm. Sci. 64, 599–609. Sotomatsu, T., Nakagawa, Y., and Fujita, T. (1987). Quantitative structure–activity studies of benzoylphenylurea larvicides. 4. Benzoyl ortho substituent effects and molecular conformation. Pestic. Biochem. Physiol. 27, 156–164. Sotomatsu, T., Shigemura, M., Murata, Y., and Fujita, T. (1993). Octanol/ water partition coefficient of ortho-substituted aromatic solutes. J. Pharm. Sci. 82, 776–781. Southworth, G. R., Keffer, C. C., and Beauchamp, J. J. (1980). Potential and realized bioconcentration. A comparison of observed and predicted bio-concentration of azaarenes in the fathead minnow (Pimephales promelas). Environ. Sci. Technol. 14, 1529–1531. Stewart, J. J. P. (1990). “MOPAC Manual,” Version 6.0. Quantum Chemistry Program Exchange. Indiana University, Bloomington, Indiana. Taft, R. W. (1956). Separation of polar, steric, and resonance effects in reactivity. In “Steric Effects in Organic Chemistry” (M. S. Newman ed.), pp. 597–603. Wiley, New York. Takahashi, J., Kirino, O., Takayama, C., and Kamoshita, K. (1988). Studies on fungicidal activity of W-phenylcarbamates. IV Determination of the hydrophobicity of N-phenylcarbamates by highperformance liquid chromatography. J. Chromatogr. 436, 316–322. Takayama, C., Akamatsu, M., and Fujita, T. (1985). Effects of structure on 1-octanol/water partitioning behavior of aliphatic amines and ammonium ions. Quant. Struct.-Act. Relat. 4, 149–160. Terada, H. (1986). Determination of log Poct by high-performance liquid chromatography, and its application in the study of quantitative structure–activity relationships. Quant. Struct.-Act. Relat. 5, 81–88. Terada, H., Kitagawa, K., Yoshikawa, Y., and Kametani, F. (1981). Partition and ion-pair partition of 2,4-dinitrophenol, an uncoupler of oxidative phosphorylation. Chem. Pharm. Bull. 29, 7–14. Uchida, M., and Kasai, T. (1980). Adsorption and mobility of fungicidal dialkyl dithiolanylidenemalonates and their analogs in soil. J. Pestic. Sci. 5, 553–558. Uchida, M., Funayama, S., and Sugimoto, T. (1982a). Bioconcentration of fungicidal dialkyl dithiolanylidenemalonates in Orizias latipes L. J. Pestic. Sci. 7, 181–186. Uchida, M., Kurihara, N., Fujita, T., and Nakajima, M. (1974). Inhibitory effects of BHC isomers on Na  K-ATPase, yeast growth, and nerve conduction. Pestic. Biochem. Physiol. 4, 260–265. Uchida, M., Nishizawa, H., and Suzuki, T. (1982b). Hydrophobicity of bupro-fezin and flutoranil in relation to their soil adsorption and mobility in rice plants. J. Pestic. Sci. 7, 397–400. Valvani, S. C., and Yalkowsky, S. H. (1980). Solubility and partitioning in drug design. In “Physical Chemical Properties of Drugs”

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(S. H. Yalkowsky, A. A. Sinkula, and S. C. Valvani, eds.), pp. 201–229. Dekker, New York. Veith, G. D., and Broderius, S. J. (1987). Structure–toxicity relationships for industrial chemicals causing type (II) narcosis syndrome. In “QSAR in Environmental Toxicology—II” (K. L. E. Kaiser ed.), pp. 385–391. Reidel, Dordrecht. Veith, D. V., and Mekenyan, O. G. (1993). A QSAR approach for estimating the aquatic toxicity of soft electrophiles [QSAR for soft electrophiles]. Quant. Struct.-Act. Relat. 12, 349–356. Veith, G. D., Call, D. J., and Brooke, L. T. (1983). Structure–toxicity relationships for the fathead minnow, Pimephales promelas: Narcotic industrial chemicals. Can. J. Fish. Aquat. Sci. 40, 743–748. Veith, G. D., DeFoe, D. L., and Bergstedt, B. V. (1979). Measuring and estimating the bioconcentration factor of chemicals in fish. J. Fish. Res. Board Can. 36, 1040–1048. Verhaar, H. J. M., Eriksson, L., Sjoestroem, M., Schiiurmann, G., Seinen, W., and Hermens, J. L. M. (1994). . Quant. Struct.-Act. Relat. 13, 133–143. Verhaar, H. J. M., van Leeuwen, C. J., and Hermens, L. M. (1992). Classifying environmental pollutants. 1. Structure–activity relationships for prediction of aquatic toxicity. Chemosphere 25, 471–491. Verloop, A. (1983). The STERIMOL approach: Further development of the method and new application. In Pesticide Chemistry: Human Welfare and the Environment (J. Miyamoto and P. C. Kearney, eds.)Vol. 1, pp. 339–344. Pergamon, Oxford. Wang, P. S., and Lien, E. J. (1980). Effects of different buffer species on partition coefficients of drugs used in quantitative structure–activity relationships. J. Pharm. Sci. 69, 662–668. Wauchope, R. D., Buttler, T. M., Hornsby, A. G., Augustijn-Beckers, P. W. M., and Burt, J. P. (1992). The SCS/ARS/CES pesticide properties database for environmental decision-making. In Reviews of Environmental Contamination and Toxicology (G. W. Ware ed.)Vol. 123, pp. 1–25. Springer-Verlag, New York. Yamagami, C., and Fujita, T. (1995). Hydrophobicity parameter of heteroaromatic compounds derived from various partitioning systems. In “Classical and Three-Dimensional QSAR in Agrochemistry” (C. Hansch, and T. Fujita, eds.), ACS Symposium Series 606, pp. 36–47. Am. Chem. Soc, Washington, DC. Yamagami, C., Ogura, T., and Takao, N. (1990). Hydrophobicity parameters determined by reversed-phase liquid chromatography. I. Relationship between capacity factors and octanol–water partition coefficients for monosubstituted pyrazines and the related pyridines. J. Chromatogr. 514, 123–136. Yamagami, C., Takao, N., and Fujita, T. (1995). Analysis and prediction of 1-octanol/water partition coefficients of substituted diazines with substituent and structural parameters. In “QSAR and Drug Design— New Development and Applications” (T. Fujita ed.), pp. 153–183. Elsevier, Amsterdam. Yamagami, C., Yokota, M., and Takao, N. (1994). Hydrophobicity parameters determined by reversed-phase liquid chromatography. IX. Relationship between capacity factor and octanol–water partition coefficient of monosubstituted pyrimidines. Chem. Pharm. Bull. 42, 907–912. Yukawa, Y., and Tsuno, Y. (1959). Resonance effect in Hammett relationship. II. Sigma constants in electrophilic reactions and their intercorrelation. Bull. Chem. Soc. Jpn. 32, 965–971. Zakarya, D., Belkhadir, M., and Fkih-Tetouani, S. (1993). Quantitative structure–biodegradability relationships (QSBRs) using modified autocor-relation method (MAM). SAR QSAR Environ. Res. 1, 21–27.

Section IX

Public Health Regulation and Epidemiology

(c) 2011 Elsevier Inc. All Rights Reserved.

Chapter 59

Studies in Humans Wayland J. Hayes, Jr., James Bruckner, and John Doull Kansas Life Sciences Innovations, Kansas City, Kansas

In response to several requests, we have included Wayland J. Hayes’ classic chapter “Studies in Humans” from his first Handbook of Pesticide Toxicology in this edition of the handbook. We have modified it gently where necessary and have added a new section to update the developments that have taken place since it was first published in 1991. The reason for performing some studies on humans rather than limiting them to animals is that species are not identical. For a given compound, the factor of species is likely to involve a greater quantitative difference than any other factor in toxicity except dosage. Furthermore, species differences may be qualitative as well as quantitative. Phenomena of mental state or of allergic sensitivity can be studied in animals only with difficulty. Because humans are unique in their response to chemical exposure, the introduction of a biologically active chemical into commercial use without first submitting it to careful observation in a limited number of people exposed to it under conditions controlled as carefully as possible has been questioned. The only alternative is to introduce the chemical directly from the laboratory to the public under conditions that necessarily minimize control and discourage observation. Information about different aspects of the effect of a chemical on humans can be obtained in four situations: (a) cases of poisoning, (b) occupational or other routine exposure, (c) use of the compound as a drug, and (d) administration of the compound to volunteers. The value of information derived from cases, therapeutic use, and occupational medicine is generally recognized, and little discussion of the techniques required for gathering such information is needed here. In this chapter, special emphasis is placed on studies in volunteers because this kind of study is less used and less well understood. This emphasis should not detract from the fact that the scientific and moral requirement is for study in humans, not necessarily for studies in volunteers. If the requisite information can be obtained from people with occupational exposure, so much the better. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

Each of the four sources of information on toxicology in humans has its own advantages and limitations. In general, cases of poisoning constitute our only source of information on symptomatology in humans. The fact that poisoning occurred is evidence that the dose was large, but its measurement is often unsatisfactory. Very detailed information on the limited but known dosage range is available for pesticides that have been used as drugs, but only a limited number have been used that way. Use experience (whether at the relatively high dosage rates of workers or the lower dosage rates of other groups) has the advantage of involving relatively large and diverse groups of people for extended periods. It has the disadvantage that the dosage is difficult to measure. Studies of volunteers offer the great advantage that their dosage can be determined accurately. Unfortunately, such studies are usually limited to small groups and short durations. The fact that volunteers cannot be given dosages high enough to cause harm is a strict limitation, but not one that seriously interferes with the value of this source of information. This chapter is concerned with techniques, especially those applicable to volunteers. It also comments selectively on pesticides that have been studied in humans, whether in connection with cases of poisoning, use as drugs, exposure of workers, or investigations in volunteers.

59.1  Cases Everyone recognizes the need to study cases of poisoning. They are our only source of information on the exact form of illness that adequate doses of most pesticides produce in people. Usually, they are also the only accurate source of information on the dosage required to produce illness. Since there is no way to predict when poisoning will occur, its study must be opportunistic. However, the general quality of reporting has improved because facilities for measuring pesticides and their metabolites have improved. Further progress could be made if more attention were 1255

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given to collection of appropriate samples. A specimen of the formulation supposed to have caused illness should be sent to the laboratory in a completely separate container from other specimens. Grossly contaminated clothing should be shipped separately also, lest vapor of the chemical contaminate other specimens. Specimens of vomitus, stomach washings, urine, feces, blood, and organs are always appropriate. Where possible, serial samples should be obtained. In addition, samples should be selected from those materials that, on an epidemiological basis, seem to have been the cause of trouble. For example, if the clothing was contaminated, samples of it should be collected and soiled and unsoiled portions analyzed. If food appears to have been responsible, the remains of what led to illness should be collected, as should samples of the ingredients of that particular food. Always sample selectively and in reference to the cases of illness.

59.2  Medical use A total of 64 pesticides have been or are used currently as drugs. After some of the older materials such as arsenic and phosphorus had been used widely, it was recognized that their toxicity far outweighed any therapeutic value that they might have and they were gradually withdrawn from medical use. There were 18 pesticides used orally and 8 others that were applied dermally for control of external parasites in 1990 but only the insect repellants and acaracides are used today (American Medical Association, 1986).

59.3  Use experience As suggested in the introduction to this chapter, the term “use experience” can be thought of as including total experience in the use of a compound. It would include any clinical findings and the results of measuring residues in people’s environments and in their tissues and excreta. It would even include such useful but negative information as the observation that a compound has been produced at a measured rate for many years, but is not known to be stored in tissues or to be the cause of any illness in humans. However, in this chapter we shall discuss only the portion of use experience that involves relatively high dosages and leads to some quantitative information. For practical purposes this limits the subject to studies in workers. The major disadvantage in using workers for tests is the difficulty in regulating or even measuring their dosage. However, this is not a handicap in many biochemical studies, especially when the results can be related to measurement of the compound or its metabolite(s) in blood or excreta. Furthermore, if the compound or its metabolite(s) can be measured, the results of the study can be used

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directly in interpreting the safety of other persons exposed under different circumstances that lead to the same, or different, blood levels or excretory levels. If the dosage is unknown, it may not be possible to determine what proportion of the compound or metabolite(s) is actually detected. Even so, it may be possible to show that the results are exposure related by comparing them with results of environmental measurements. Often, the ideal scheme is to use volunteers to relate carefully measured dosages to the kinetics of storage and excretion, and then to carry out other studies, especially very prolonged ones, in workers. Certainly, the exposures of workers often last much longer than any that would be practicable in volunteers. At least 142 pesticides have been studied in exposed workers from an epidemiological, clinical, or chemical standpoint. This number includes studies of mixers and sprayers as well as the inhabitants of the houses they sprayed in connection with village trials of the effectiveness and safety of compounds for the control of vectorborne diseases (mainly malaria). The number does not include studies of poisoning resulting from gross accidents even though the accidents were occupational in origin.

59.4  Volunteers 59.4.1  Introduction The use of volunteers to study pesticides is less understood and, therefore, more controversial than the study of (a) cases of poisoning by these compounds, (b) the therapeutic use of drugs (including those that happen to be or that become pesticides), or (c) workers exposed to pesticides. The most important advance that has been made in our times regarding planned tests in humans is the recognition that these tests constitute established medical practice, at least for evaluating the effectiveness of drugs. Testing, as such, is not new. It has been carried on to one degree or another since organized medicine came into existence. As early as 1803, Percival outlined the ethical principles for testing new remedies in individual patients. However, occasional systematic tests were carried on at least as early as 1721 when Dr. Richard Mead, physician to King George I, inoculated with smallpox (variolation) six condemned felons who had submitted on condition of being pardoned. The King’s pardon was the result of a request by Caroline, Princess of Wales, who had learned of variolation from Lady Mary Wortley Montague, who had observed it in the Near East. Five of the volunteers contracted mild disease; the sixth, who concealed that he had had smallpox previously, was not infected, but he too escaped hanging. A young woman was infected by inhalation; she became much sicker than the men but survived. In his medical account of the experiment, Mead (1747)

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attributed the method of inoculation to the Greeks and the method of inhalation to the Chinese and indicated the superiority of inoculation. It is also recorded that, following the successful outcome of the test on the six men, the King agreed to the variolation of the royal children but refused to have it himself. Other early studies included that of James Lind (begun in 1747 and discussed subsequently) and that of Dr. Louis (1835), who investigated the effect of bleeding on the survival of patients with pneumonia. He found that the longer bleeding was delayed after the onset of illness, the higher the proportion of patients who survived. He concluded quite rightly that if bleeding had not been carried out at all, the results would have been even better. Most people in the medical profession are familiar with the fact that tests on the efficacy of drugs are routine. It is also generally known within the profession that such studies are done with improved design and increasing skill. If the frequency of these tests is not sufficient evidence that they constitute established medical practice, then certainly this status was achieved in October 1962, when the Drug Amendments Act of 1962, which amends the Federal Food, Drug, and Cosmetic Act (21 U.S.C. 321 et seq.), was passed in the United States, requiring that the effectiveness of all new drugs be tested in humans before the drug can be licensed for sale. However, the law requiring that the effectiveness of drugs be tested in humans does not require specifically that their safety be tested in this way. It does demand that safety be tested in animals and that any untoward effects in humans be reported promptly. If injury of sufficient importance is demonstrated, license for sale of the drug will be withheld or withdrawn, depending on whether or not the drug has actually reached the market. The fact that testing of effectiveness in humans is required directly but testing of safety in humans is required only indirectly could be a matter of semantics, but the difference probably reflects a certain general ambivalence in present thinking about tests in humans. This ambivalence may be explained on the one hand by an ever-present awareness that such tests are needed and on the other hand by a combination of fear that the tests will be misused and ignorance of the manner in which these tests are carried out under proper scientific supervision. The ambivalence can be relieved by increasing attention to the ethical necessity of testing chemicals purposefully, scientifically, and safely in humans rather than leaving the matter to accident.

59.4.1.1  Kinds of Studies A number of authors have recognized tacitly (Medical Research Council, 1964) or expressly (Académie Nationale de Médecine, 1952; Hayes, 1965) that there are different kinds of studies in volunteers. Although different terms

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may be used, all authors agree that the basic distinction is between (a) studies of diagnostic procedures or therapy and (b) studies of physiology in the broadest sense. (a)  Diagnostic or Therapeutic Subjects for the first kind of study are usually sick at the time of the study. They or their representatives freely consent to a calculated risk in undertaking the study, but they do so in the hope that the diagnostic procedure or treatment will be of benefit to their recovery. Such studies are generally carried out according to a sophisticated design, frequently by the double-blind technique. An early example of a controlled study of this kind was that by which James Lind (1753) compared the value of oranges and lemons, vinegar, sea water, cider, and a variety of other preparations as cures for scurvy among 12 patients he selected for the experiment. However, it is generally recognized that if the disease being diagnosed or treated is sufficiently rare, such studies must be carried out with a less formal plan, although with no less skill. For example, a promising new treatment for an unusual form of poisoning must be used when the opportunity arises or it will not be used at all. In the case of patients who are unable to give consent because of their age or medical condition, consent should be obtained from the parents or next of kin, just as for any diagnostic procedure or therapy for such patients. It is agreed (Freund, 1965) that parents or guardians may consent for a child (or other legal incompetent) with respect to acts for the welfare or benefit of the child (or other legal incompetent). (b)  Physiological Studies Subjects for the second kind of study may be sick or well. In either event the state of their health is not in itself critical to the study, and the subjects have no reason to suppose they will be healthier because of the test. They are often paid or given some other benefit in exchange for their service and the calculated risk to which they freely consent. (Minors and mental incompetents should not be accepted as volunteers, or should be accepted only with caution and with the consent of their guardians.) In some jurisdictions, only a parent (or other person) who has been legally appointed as a guardian can give consent for the participation of a child in a study from which the child receives no therapeutic benefit. However, this question will rarely, if ever, arise in connection with the study of pesticides because children will rarely participate unless they have been poisoned accidentally and the study involves only therapy. Regardless of the kind of study, insofar as possible, children should be informed and should give their own consent. Here again, the formality of the plan of study has to depend on the situation. Most studies involve preselected groups of subjects, including one or more control groups.

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On the other hand, Beaumont (1833) had no possibility of using controls or of checking his observations in a number of preselected subjects, for there was only one Alexis St. Martin. Although the separation of tests in humans into the two kinds just described is generally valid, it implies a certain defect of logic already hinted at. To be sure, tests of safety or pharmacological action are not themselves therapeutic or diagnostic, even if they involve a drug. However, patients selected as controls in therapeutic studies receive no benefit to their own health. The contradiction lies in the fact that certain authors have contended that the only permissible tests are those that potentially benefit the health of every individual who participates in them. Strict application of this concept would rob therapeutic tests of much of their clinical value and scientific worth. It is true that the rewards to the participant in the two kinds of studies described above are often different, requiring different administrative approaches. However, the aim and contribution of both kinds of studies is the ultimate improvement of human health. The basic similarity in risk of the two kinds of studies should be emphasized also. There is no way to exclude potential danger from any study. The simplest procedure, such as venipuncture, occasionally causes serious injury, even when proper precautions are observed. However, this is no more true of studies involving pharmacology or safety than of those involving diagnosis or therapy. The probable risk must be judged for each test individually.

59.4.2  Legal and Ethical Considerations 59.4.2.1  Codes and Guidelines (a)  Codes Table 59.1 lists some of the codes and guidelines for regulation of one or more kinds of medical study in humans. The subject is not new; the first known guideline was dated 1803, as discussed in Section 59.4.2.3. The Nuremberg Code was developed against a background of inhumane and unscientific experiments carried out under Nazi political and military pressure. However, earlier codes, including the German codes of 1901 and 1931, as well as later codes were the result of ethical self-regulation of the medical profession. The content of all the codes, including the Nuremberg Code, is notable for its consistency. There are, of course, those who think that the more detailed and rigid the code, the more likely it is to accomplish its objectives. However, it seems unlikely that additional rules will provide any real benefit to anyone if the following three requirements specified by the American Medical Association (1946) are complied with scrupulously. (a) There must be voluntary consent of the person on whom the investigation is to be performed. (b) The danger of each study must be previously investigated by animal

Table 59.1  Some Codes and Guidelines for Medical Study in Humans Name (subject)

Reference

New remedies and new methods of chirurgical treatment

Percival (1803)

Instructions to directors of clinics, etc.

Ministry of Religious, Educational and Medical Affairs (1901)

Guidelines on innovative therapy and scientific experimentation

Minister of the Interior (1931)

Supplementary report of the Judicial Council

American Medical Association (1946)

Nuremberg Code

Nuremberg Military Tribunal (1949)

The moral limits of medical research and treatment

Pope Pius XII (1952)

Report on human experimentation

Public Health Council of the Netherlands (1957)

Responsibility in investigations on human subjects

Medical Research Council (1964)

Some considerations in the use of human subjects in safety evaluation of pesticides and food chemicals

National Academy of Sciences (1965)

Declaration of Helsinki (Helsinki I)

World Medical Association (1964)

Declaration of Helsinki, revised at Tokyo (Helsinki II)

World Medical Association (1976)

Proposed International Guidelines

WHO/CIOMS (1982)

experimentation. (c) The investigation must be performed under proper medical protection and management. In the last analysis, the conduct of tests in humans depends on the good sense, integrity, and freedom of the participating scientists, rather than on an elaborate code of ethics. In spite of this, it is important for historical purposes to repeat the ten points laid down by the Nuremberg Military Tribunal (1949). 1. Voluntary consent of the subject is absolutely essential. Consent must be based on knowledge and understanding of the elements of the study and awareness of possible consequences. The duty of ascertaining the quality of consent rests on the individual scientist and cannot be delegated. [The original document, which was prepared as a basis for much of the Nuremberg Code, contained provisions for valid consent of mentally sick persons to be obtained from the next of kin or legal guardian and

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also from the patient, whenever possible. These special provisions were omitted from the Code, probably because this kind of consent did not apply in the specific cases under trial (Alexander, 1966).]   2. The test should seek some benefit to society unobtainable by any other method.   3. The investigation should be designed and based on prior animal study and the natural history of the disease or problem, and other data, so that anticipated results may justify the action taken.   4. It should be so conducted as to avoid unnecessary physical and mental suffering.   5. No test should be undertaken where there is reason to believe that death or lasting disability will occur, except perhaps where the investigator may serve as his own subject.   6. The degree of risk should never exceed that which the importance of the problem warrants.   7. There should be preparation and adequate facilities to protect the subject against even remote possibility of injury, disability, or death.   8. Only scientifically qualified persons, exercising a high degree of skill and care, should conduct investigations on human beings.   9. The subject should be permitted to end the test whenever he reaches a mental or physical state in which its continuation seems to him impossible. 10. The investigator must be prepared to end the test if he has reason to believe that its continuation is likely to result in injury, lasting disability, or death. Any promulgation of a code or any discussion of principles for tests in humans is evidence for acceptance of such tests under certain conditions. It is therefore significant that in several countries codes or discussions of principles, which are basically similar and consistent, have been prepared by important persons or institutions, including Pope Pius XII (1952), the Medical Research Council (1964), the Public Health Council of the Netherlands (1957), the National Academy of Sciences (1965), the American Medical Association (1946), the U.S. Public Health Service (Stewart, l966a), and the World Medical Association (1964, 1976). Approval of tests in humans is most clearly implicit in those codes of practice that set forth administrative regulations for conduct of such tests in a particular institution or in laws that require such tests. Examples are discussed in the next sections. (b)  Guidelines Apparently the only new aspects of the codes are guidelines for their interpretation. The distinction between codes and guidelines in this instance is interesting. Those who have been active for several years in formulating the guidelines accept the internationally recognized codes without

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question but have been concerned with research involving human subjects as a problem of developing countries. The guidelines proposed will offer some countries nothing that is not already in force in one form or another. They have been framed with special reference to the requirements of developing countries and elaborated in the light of replies to a questionnaire received from 45 national health administrations and 91 medical faculties of countries in which medical research involving human subjects is as yet undertaken on a limited scale and/or in the absence of explicit national criteria for protecting such subjects from involuntary abuse. The replies were received from a total of 60 developing countries. The proposed guidelines were prepared jointly by the World Health Organization and the Council for International Organizations of Medical Sciences (WHO/ CIOMS) (1982) because of the prospect that more applied biomedical research would henceforth be undertaken in developing countries. Persons developing the guidelines have concluded that some considerations have special relevance to work undertaken in developing countries. As one might expect, most of the considerations reviewed in connection with the guidelines are concerns of all humans, with no special relevance to the degree of national development. The only considerations that clearly have special relevance to developing countries involve either external sponsorship or the present lack of relevant legislation in some countries. Complications that may be associated with external sponsorship are several. The investigation may subserve external rather than local interest. Foreign investigators and sponsors may not possess adequate insight into local mores, customs, and legal systems. By the same token, failure to understand the community as an integral whole may lead to unintentional neglect of the people’s expectations. The absence of any long-term commitment to subjects involved in the research and withdrawal of outposted personnel on completion of their task may result in local disillusionment. Lack of accountability may deprive subjects of any form of compensation for incidental injury. Diversion of highly trained local medical personnel – a scarce and valuable resource – to research activities may constitute an inordinate strain on community resources. The research activity is most readily justified when the objectives are clearly attuned to important issues of local relevance, when service or training can be offered through an established local medical institution, or when both conditions apply.

59.4.2.2  A Neglected Ethical Principle A special aspect of the ethical objective of benefit to the community involves the question of whether it is ethical to introduce a new chemical, regardless of its use, without assuring that it will be studied in humans either before it is introduced or as an integral part of a plan for its

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introduction. The question is not whether every compound ought to be studied in volunteers but whether every compound ought to be studied systematically. The alternative of some systematic studies is to release the compound without a plan for monitoring any unfavorable effects it may have. This question was resolved long ago in the case of drugs. Studies of the safety of drugs often are not required in the same way that studies of the effectiveness are required. The ambivalence of our thinking reaches full expression in connection with nontherapeutic xenobiotics. It must be recognized that there is considerable reluctance to test the safety of these materials in humans. This is true even though the ethical imperative to make such studies has been expressed for years. For example, Keplinger (1963) wrote, “the toxicologist has a moral obligation to evaluate the safety of chemicals in the best possible manner” and also “it can be concluded that the proper testing on man by qualified scientists should be considered an integral phase of a toxicology investigation.” Hayes (1968) expressed the matter as follows: “little attention seems to have been given to the question of whether it is ethical to release a new chemical for general use before it has been tested under controlled conditions in a limited population.” Paget (1970) wrote If such chemicals are to be brought into use and widely distributed the alternative is not whether human beings should be used in toxicity experiments or not but rather whether human beings should be used in small numbers in controlled toxicity experiments or whether large populations should be exposed to chemicals, knowledge about which is confined to their effects on other animal species. In these circumstances, it is clear that the rational course, before any new chemical, whether drug, pesticide, or food additive, is added to the human environment, is for cautious, closely controlled and observed trials to be carried out in groups of human beings.

This view was paraphrased and endorsed by an international group of experts sponsored jointly by the World Health Organization (WHO) and the United National Environmental Programme (1978).

59.4.2.3  Regulation of Clinical Research and Investigations (a)  Group Review Group review is an old idea that has received increased emphasis during the last 20 years. Percival (1803) made no claim to originality when he wrote, regarding new remedies and new methods of chirurgical treatment, in the accomplishment of this salutary purpose, the gentlemen of the faculty should be scrupulously and conscientiously governed by sound reason, just analogy, and well authenticated facts. And no such trials should be instituted without a previous consultation of the physicians or surgeons according to the nature of the case.

In more recent times, the idea of consultation has been formalized and extended in the form of ethical review committees. Apparently the first requirement for such a committee was issued on November 17, 1953; it applied only to intramural research at the then new Clinical Center of the National Institutes of Health in Bethesda, Maryland. The pattern was extended to extramural research financed by the U.S. Public Health Service by a resolution of the National Advisory Health Council on December 3, 1965, and formalized by a policy decision issued by the Surgeon General on February 8, 1966 (Stewart, l966a), which contained the following statement. No new, renewal, or continuation research or research training grant in support of clinical research and investigation involving human beings shall be awarded by the Public Health Service unless the grantee has indicated in the application the manner in which the grantee institution will provide prior review of the judgment of the principal investigator or program director by a committee of his institutional associates. This review should assure an independent determination (1) of the rights and welfare of the individual or individuals involved, (2) of the appropriateness of the methods used to secure informed consent, and (3) of the risks and potential medical benefits of the investigation. A description of the committee of the associates who will provide the review shall be included in the application.

It was pointed out that such a “committee of institutional associates” would need to be (a) acquainted with the investigator under review, (b) free to assess the investigator’s judgment without placing in jeopardy their own goals, (c) sufficiently mature and competent to make the necessary assessment, and (d) composed in part of members drawn from different disciplines or interests that do not overlap those of the investigator under review. The policy was implemented by a number of administrative issuances (U.S. Public Health Service, Division of Research Grants, PPO No. 129, February 8, April 7, July 1, 1966) that established the coverage, form, and timing of the assurance each institution must furnish before it may receive additional grants or extensions of existing grants. It is required that the grantee make and keep written records of group reviews and decisions and obtain and keep documentary evidence of informed consent. The policy was also extended to intramural and contract research carried out by the Public Health Service (Stewart, l966b). The requirement for institutional review boards has been amplified and has become firmly established in regulations. In the United States, there are separate regulations for different departments and even for different portions of the same department of the government. The regulations for the Public Health Service covering in-house research and that carried out under grants and contracts may be found in 45 CFR 46. Similar regulations for the Food and Drug Administration, now a part of the Public Health

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Service, are in 21 CRF 56; this regulation covers research sponsored by manufacturers of drugs and carried out for the purpose of obtaining a marketing permit. A detail of the regulation [21 CFR 56.103(b)] states that the agency may refuse to consider data gathered without compliance. Presumably, this extends coverage to studies carried out for this purpose in any other country. A separate set of regulations (21 CFR 50) covers those instances in which research related to the Food and Drug Administration is to be carried out in prisoner volunteers, including those held by state or local prisons. The Department of Energy not only requires institutional review boards but also specifically prohibits, in agreements with volunteers, any exculpatory clause that waives or appears to waive any of the volunteers’ legal rights, including any release of the institution or its agents from liability for negligence (10 CFR 745). Regulations of the Bureau of Prisons take a somewhat different form in that its research committees are all associated with the Bureau, either at the institutional, regional office, or central office level (32 CFR 219). The same regulations limit to soft drinks and snacks the tangible incentives that may be offered to federal prisoners, but they do not restrict incentives based on relief of boredom, on patriotism, or on contribution to medical science, perhaps the main motives for the voluntary participation of prisoners. Of course, the regulations of different departments determine that if a study in prison volunteers is funded by another federal agency, each study will have to be reviewed by at least two boards, one for the institution carrying out the study and one within the Bureau of Prisons. Regulations for volunteers within the Department of Defense are contained in 32 CFR 219, where some details are covered by reference to the Public Health Service regulations (45 CFR 46). Review boards related to the Department of Defense differ from others in having their authority limited to scientific and ethical review with authority to actually initiate or halt studies reserved to the command where each study is done or to a superior command. Discussion of institutional research committees has spread worldwide. Recent papers on the subject have appeared in German, French, Spanish, Japanese, and Russian, and they have appeared in English in the United States, the United Kingdom, Canada, Scandinavia, and India. During recent years, about 35 papers regarding human experimentation have been indexed each year in Index Medicus. There is no possibility of reviewing all of this literature in this chapter. Much of it is repetitious. Not a few papers, while proposing no fundamental change, are openly critical of the bureaucracy that seems to be an inevitable part of institutional review committees. An editorial (Anonymous, 1984) spoke of “regulated progress or damned interference.” Osborne (1983) spoke of “the wrong balance of administration and ethics.” Pattullo (1982) spoke of “a plan that would make all scientific

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inquiry into human affairs uniquely subject to federal control.” He also spoke of “those who make applied ethics their profession.” In summary, no one doubts the value of group review. The advantage generally emphasized at this time is its ability to prevent abuses, especially failure to obtain informed consent. However, insofar as a group consists of biomedical specialists, their review may lead to technical improvements in a plan of study. Essentially no recognition has been given to the fact that, like all other human institutions, group review itself is subject to abuses, mainly bureaucracy. So much time may be spent in debating trivial matters that the real merit of a candidate study is lost sight of and the study may fail, not through any significant fault but through a bureaucratic fear of taking a positive stand. Any discussion of the practical realization of codes of conduct must distinguish between laws that support universally accepted ethical principles and restrictive regulations that may be issued under these laws. Those who issue such regulations should not always be blamed for them; some elected politicians have criticized studies in humans in order to create an image of themselves as defenders of the oppressed and the disadvantaged. Under these circumstances, civil servants or academic bureaucrats find it difficult to avoid the temptation to make all regulations as stringent as possible.

59.4.2.4  A Law Permitting Tests in Humans A law that permits tests in humans is the Armed Forces Act of August 10, 1956 (10 USC 4503), representing the revision, codification, and enactment of pre-existing laws relating to the military services. Army Regulations No. 70-25 (Use of Volunteers as Subjects of Research, Washington, D.C., March 26, 1962) cites the Armed Forces Act as its basis. Army Regulations No. 70-25 refers only to tests on humans that pertain to military research and specifically excludes “… investigations involving the basic disease process or new treatment procedures conducted by the Army Medical Service for the benefit of patients.” With the exception of any specific reference to prior studies in animals, the regulations restate the ten points laid down by the Nuremberg Military Tribunal (see Section 59.4.2.1). In addition, the regulations impose a number of other limitations. Prisoners of war will not be used under any circumstances. A physician approved by the Surgeon General of the Army will be responsible for the care of volunteers. (The physician may or may not be project leader, but he will have authority to terminate the study at any time.) Appropriate consultants will be available at all times. No studies will be done prior to approval of each protocol by the Surgeon General of the Army and the Chief of Research and Development; for certain classes of studies, approval of the Secretary of the Army is required.

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59.4.2.5  Legal and Other Requirements for Tests in Humans It is clear that all countries that are successful in developing new drugs find some way to test them in humans. (a)  United States Apparently the only law clearly requiring tests in humans as a condition for the licensing of new drugs is the Drug Amendments Act of 1962 mentioned in Section 59.4.1. The law provides, among other things, that a license will be refused for any new drug if “there is a lack of substantial evidence that the drug will have the effect it purports or is represented to have under the conditions of use prescribed, recommended, or suggested in the proposed labeling thereof.” According to the law the term “substantial evidence” means evidence consisting of adequate and wellcontrolled investigations, including clinical investigations, by experts qualified by scientific training and experience to evaluate the effectiveness of the drug involved, on the basis of which it could fairly and responsibly be concluded by such experts that the drug will have the effect it purports or is represented to have under the conditions of use prescribed, recommended, or suggested in the labeling or proposed labeling thereof. The same law also provides that a license will be refused for any new drug if the reports that are required to be submitted to the Secretary “do not include adequate tests by all methods reasonably applicable to show whether or not such drug is safe for use under the conditions prescribed, recommended, or suggested in the proposed labeling thereof.” Having provided that no new drug will be licensed without prior tests in humans for efficacy, the act conditions all such tests upon the manufacturer, or the sponsor of the investigation, requiring that experts using such drugs for investigational purposes certify to such manufacturer or sponsor that they will inform any human beings to whom such drugs, or any controls used in connection therewith, are being administered, or their representatives, that such drugs are being used for investigational purposes and will obtain the consent of such human beings or their representatives, except where they deem it not feasible or, in their professional judgment, contrary to the best interests of such human beings. It is interesting that the law requires the consent of persons who serve as controls as well as those who take or receive candidate drugs. The fact that provision is made for professional judgment in the matter of consent is noteworthy. As pointed out by Cady (1965), the discussion of this matter of consent by the legislators who passed the law (Congressional Record, 1962) will probably be considered if the question of professional judgment in connection with consent should ever be involved in an action in federal court.

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(b)  France The French law on new drugs (deGaulle et al., 1959), as explained in its preamble, was made necessary by the Stalinon catastrophe. This circumstance probably explains the emphasis accorded to safety in Article L. 601 of the law. The law does not specifically require tests in humans, but clinical trials are required by a regulation (Article R. 5 119) issued under the law. Article R. 5119 outlines a wide range of reports that must be submitted in the form of two dossiers in connection with each application for a registration or visa. Among these reports, the one on clinical trials must specify any necessary contraindications and define the conditions of use. This regulation implies that those making clinical trials will be alert to untoward reactions during the course of the work. However, the separate paragraphs of Article R. 5119 dealing with toxicology and proof of safety do not mention tests in humans. In fact, the requirement (Article R. 5.20-4) that the clinical experts be furnished reports of all trials of toxicity before beginning clinical trials suggests that the toxicity tests are limited to animals. An important feature of the French law is the requirement, set forth in Article L. 605, that compliance of a candidate drug with the requirements of Article L. 601 must be verified by experts chosen by the manufacturer from a list established by the Minister of Public Health and Population. Article R. 5122 sets the term of office of the experts and establishes the composition of a commission to nominate them. For further details of the workings of the French law see Guillot (1964). It is interesting that the French law and regulations apparently do not require, or even mention, consent by patients taking part in clinical trials. (c)  United Kingdom Apparently no law in the United Kingdom requires or specifically regulates tests in humans. However, somewhat the same goal is accomplished by the Committee on Safety of Drugs (1965) through a “Memorandum to manufacturers and other persons developing or proposing to market a drug in the United Kingdom.” In one sense, this brief Memorandum is broader that the American or French laws, for it refers to “pharmacological studies on healthy human volunteers” as a part of the preclinical investigation of a new drug. The Memorandum goes on to state that the results of clinical trials should include “(a) human pharmacological findings; (b) therapeutic activity; (c) any hazards, contra-indications, side effects, and necessary precautions.” Although the Committee on Safety of Drugs is not a part of the Ministry of Health, it is financed by the Ministry. The Committee would seem to be in an excellent position to sponsor legislation that would give protection from tort claims to the medical scientist who carefully and

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properly conducts medical research on human beings in accordance with prescribed standards. Under such legislation, the acts of the scientists, within specifically prescribed limits, would be those of the government, and “sovereign immunity” would be waived in cases where unforeseen or unpredictable harm results to a subject. (d)  Sweden Rydin (1965) and Lonngren (1965) have discussed toxicity testing of drugs in Sweden and its relationship to laws concerning drugs in Sweden. New drugs must be registered before use. The National Pharmaceutical Laboratory is in charge of the majority of the control work. The Council on Drug Acceptance is the consulting organ, and the National Board of Health determines whether or not a drug can be accepted for registration. Clinical trials include pharmacological and toxicological as well as therapeutic considerations. Emphasis is placed on individual handling of each application for registration. (e)  Discussion American regulations place more restrictions on the investigator of new drugs than apparently are imposed by medical ethics and custom in some other countries where greater emphasis is given to the discretion of the physician and less to the consent of the patient. However, this relative disregard for the consent of patients subjected to new drugs may be linked with customs forbidding all tests in humans that do not contribute, at least potentially, to the health of the subject. As already pointed out, strict insistence on direct medical benefit to the patient rules out not only studies devoted primarily to pharmacology or the safety of chemicals including drugs, but also the use of placebos in tests of efficacy. Thus, truly scientific testing of drug efficacy would be prevented by a requirement that all tests contribute, at least potentially, to the health of every subject.

59.4.2.6  Court Action (a)  Malpractice There have been very few court cases involving tests in humans as discussed in this chapter. Of course, negligence (malpractice) may be a basis for suit at any time. Since the generally accepted landmark case of Slater v. Baker in 1767, English and American courts generally have held that human experimentation is at the peril of those who conduct the tests. In the case of Slater v. Baker, tried in English courts (King’s Bench) under the jury system, it was established (presumably on the assumption that the physician deviated from the then current method of treating a fractured tibia) that the physician lacked proper skill and failed to disclose his intention or to obtain the patient’s consent in using a previously untested procedure. The case

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established a precedent that has been followed in English and American law and, in various degree, in the laws of other countries. In any case in which negligence is alleged, the true issue of investigation does not arise. Should a claim arise in the course of a planned test in humans, the best defense is proof that the subject assumed the risk knowingly and voluntarily (Monthly Summary from the Office of General Counsel, U.S. Department of Health, Education, and Welfare, December, 1959). Certainly, the mere fact of investigation is no longer considered evidence of negligence. As Cady (1965) has pointed out, the question is not whether a study is made, but how well it is made. In France, a physician found guilty of malpractice may be subject to fines or a prison sentence under Articles 319, 320, and R.4, fourth paragraph on the penal code, or to damages under Article 1382 of the civil code. (b)  A Case Involving Scientific Tests in Humans The only case brought to trial involving scientific tests in humans was the matter of Hyman v. Jewish Hospital (15 NY 2d 3 17), in which Hyman, a member of the board of directors of the Jewish Chronic Disease Hospital, sought the right to examine the records of 22 patients who had taken part in a test. The petition was granted by the trial court, reversed by the lower state appeals court, but finally upheld by the Court of Appeals of New York, the highest court in that state. The final decision ruled that the director was entitled to examine the medical records in connection with the study because of the possibility of liability of the hospital, although no opinion whatever was expressed on whether such liability would indeed exist. The court contest attracted interest in the case from the New York State Department of Education, and the professional conduct of the two physicians primarily concerned with the study was reviewed by the Regents of the University of the State of New York, acting under their responsibility for licensing the medical profession. The findings of fact in the case were published by Freund (1965) and discussed at length by Langer (1966); Langer has also reproduced the body of the Regents’ decision. In brief, the study involved subcutaneous injection of living cancer cells into 19 very debilitated patients suffering from nonmalignant chronic disease, and into three equally debilitated control patients who had cancer. The patients were told that a cell suspension would be injected as a test for immunity or response. The patients were also told that within a few days a lump would form at the site of injection but that it would gradually disappear after a few weeks. This is exactly what happened. The record indicates that all the patients approached agreed to the injection and, further, that none suffered any ill effects other than the transient discomfort of the injection and the nodule it produced. The study showed that the sick and debilitated patients with nonmalignant disease rejected cancer

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transplants as promptly (6–8 weeks) as healthy volunteers who had been tested earlier. The test also confirmed that cancer patients are deficient in their ability to reject cancer transplants, with the rejection requiring 6 weeks to several months. The court noted that these findings opened up the possibility of stimulating defense against cancer, either before onset or perhaps even later once it had taken hold. In spite of these considerations, the Regents found the two physicians guilty of “unprofessional conduct” and of “fraud and deceit in the practice of medicine” because the patients had not been told that living cancer cells were to be injected and, therefore, the patients could not give informed consent. The physicians’ medical licenses were suspended for a year, but the sentence was stayed, and the men continued to practice while on probation for 1 year. The Regents held that (a) it is the volunteer, and not the physician, who has the right to decide what factors are or are not relevant to consent – hence all details must be revealed – and (b) when acting as an investigator, the physician has no claim in the doctor–patient relationship that in a therapeutic situation would give him the generally acknowledged right to withhold information if he judged it in the best interest of the patient. The desirability of obtaining consent in writing was emphasized. Perhaps the most important aspect of the entire proceedings was that the legality and propriety of the study itself were never challenged; only the failure to obtain informed consent was condemned. This is a strong indication of public support for medical research. Further evidence may be found in the informed and constructive comments on the case appearing in some parts of the lay press (Carley, 1966), as well as in some scientific journals (Freund, 1965; Halpern, 1966; Langer, 1966).

59.4.2.7  Consent Informed consent for a particular project ought to be based on knowledge of need for the research, as well as on knowledge of real and potential hazards. A person’s consent will be more complete if it is based on a conviction that participation is contributing to the welfare of self and others. Since the vast majority of investigators have the welfare of their subjects uppermost in their minds, it seems likely that the greatest contribution of the concept of consent is not in protecting subjects – although this is a real contribution – but rather in providing a device for justifying studies that do not offer direct benefit to the participant. If informed consent is the yardstick by which adequate protection of the subject is measured, there is no logical difficulty in undertaking studies of the safety of drugs and a wide range of other chemicals. All of these tests are excluded if one insists that every study must be of potential value to every participant. Thus, the concept of consent offers an intellectual basis for a truly scientific study of drugs and other

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chemicals, whereas insistence on personal benefits fails to offer the same intellectual basis and, if adhered to, rules out the use of placebos and double-blind study. It is argued by some that only persons trained in medicine or closely related disciplines are capable of really understanding an investigation. Although this statement contains an element of truth, it is basically misleading. Any adult of normal intelligence can be instructed to understand the most drastic potential hazards of an investigation, despite a lack of comprehension of experimental details or the probability of injury. A release from liability for “any injury, fatal or otherwise, that may result” is easily understood. It is also easily seen that a release from liability for “any results that may occur” carries the same implications (see Section 59.4.3.5). A great many studies have been completed by volunteers who signed such releases as a condition of participation. Thus, an intelligent adult is capable of informed consent. These considerations cover most situations involving studies of pesticides and other industrial chemicals in volunteers, inasmuch as nearly all these studies involve adults capable of both understanding and granting legal consent. A minor exception may involve the use of new drugs in the treatment of persons who have been poisoned but who, because of age, unconsciousness, or some other condition, are incapable of giving legal consent. However, because of the clear need of the patient for medical assistance, this small number of cases falls naturally within the ordinary patient–physician relationship and has offered no practical problem of consent. It rarely happens that there is any need to carry out nontherapeutic tests of chemicals in children or in persons who are mentally disturbed or incompetent. However, there is a very real need to use such subjects in connection with studies of nutrition, pediatric disease, psychiatric disease, and some other important conditions. The question of consent in these instances is beyond the scope of this volume but information on the subject may be found in several reviews (American Academy of Arts and Sciences, 1969; Ladimer and Newman, 1963; National Academy of Sciences, 1965, 1967; Stumpf, 1966, 1967).

59.4.3  Design of Studies 59.4.3.1  Selection of Parameters Tests in humans should be carried out only after whatever tests may be possible in animals have revealed one or more parameters suitable for study in humans. Even the most superficial study in animals often will reveal which organ system is involved predominantly in toxic action. One may then determine by further study which of the functional tests appropriate for exploring this organ system in humans is best able to measure the alteration produced in animals.

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Thus, an index test may be found. Emphasis should be placed on the appropriateness and thoroughness of the studies, not on the use of large numbers of animals. Every effort must be made to avoid producing in humans any clinical illness or any irreversible subclinical change. Therefore, in preliminary animal studies, emphasis must be placed on finding tests to detect the earliest mildest reversible change that indicates action of the chemical under study. A valuable discussion of tests of the various organ systems of humans is given in a two-volume book entitled Animal and Clinical Pharmacologic Techniques in Drug Evaluation, edited by Nodine and Siegler (1964) and by Siegler and Moyer (1967). With few exceptions, research has failed to provide a test for any effect of organic phosphorus or carbamate insecticides that is more sensitive than a test for depression of blood cholinesterase activity. This is true if the comparison is made with alterations of conditioned reflexes (Letavet, 1961) or other behavioral change (Carpenter et al., 1961). Therefore, any study of an organic phosphorus or carbamate insecticide in humans ought to include the measurement of plasma and red cell cholinesterase as index tests. The recognized exceptions of the generalization about the sensitivity of cholinesterase include the inhibition of aliesterases by low dosages of some organic phosphorus pesticides and the neurotoxicity of one organic phosphorus compound that is not a pesticide. The implication of these two kinds of exceptions is entirely different. Some organic phosphorus compounds (e.g., EPN, TOCP, dioxathion, carbophenothion, fenthion, coumaphos, ronnel, tributyl phosphorotrithioite, and schradan) inhibit aliesterases (diethylsuccinase and tributyrinase) at substantially lower daily dosage levels than they require to inhibit cholinesterases to the same degree. Other organic phosphorus compounds (e.g., parathion) inhibit aliesterases and cholinesterases at about the same dosage levels (Su et al., 1971). Compounds that inhibit aliesterases much more efficiently than they inhibit cholinesterases are likely to potentiate the toxicity of malathion and other compounds whose toxicity is highly dependent on their rate of detoxification. Compounds that inhibit aliesterases and cholinesterases to about the same degree are likely to be additive in their toxicity. Any study of an organic phosphorus compound or carbamate in humans offers an opportunity to learn its ability to inhibit aliesterases, but this measurement is less important than measurement of cholinesterases. All potential insecticides that produce neurotoxi­ city (delayed irreversible paralysis) also produce cholinesterase inhibition. Measurement of cholinesterase activity used as an index test in any investigation in humans concerned with an organic phosphorus or carbamate compound may give warning before a neurotoxic dosage is reached. However, at least one organic phosphorus compound (tri-p-ethylphenyl phosphate) that is not insecticidal and

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does not inhibit cholinesterase is neurotoxic in chickens (Cavanagh et al., 1961). Clearly, the possibility that an organic phosphorus compound is neurotoxic should be thoroughly studied, and compounds that are clearly positive should neither be tested in humans in vivo nor be introduced as pesticides. Furthermore, the activity of neuropathy target esterase (NTE), which is known to be associated with neurotoxicity, may be measured in human lymphocytes in vitro so that the susceptibility of people to this kind of injury may be explored harmlessly. Some functions may be measured mechanically rather than biochemically. Complex reaction time was found by Durham et al. (1965) to be affected in persons sufficiently exposed to organic phosphorus insecticides. It seems reasonable that the same would be true of airway resistance. Depth perception was affected by traces of tetraethylpyrophosphate in one eye, but not in both eyes (Upholt et al., 1956). In general, even those functions that are relatively simple to measure mechanically in humans are difficult or impossible to measure in animals, yet, like the effects just mentioned, they may be critical to the safety of workers or others. An effort should be made to measure the effect, if any, of a compound on the critical functions of those who may use it, whether or not there is a pharmacological reason to think it will affect those functions. For example, drugs or other chemicals to which pilots may be exposed should be tested for their effect on vision and reaction time. Any test of pharmacological action is of potential value in judging the safety of a chemical not intended for use as a drug. No action is expected under ordinary conditions of use. If one occurs, its significance must be evaluated. On the contrary, pharmacological action must be present if a compound is a drug. Thus, a drug is judged unsafe only if it leads to discomfort or dysfunction incommensurate with the benefits it produces. Whenever either drugs or other chemicals are tested in humans, an effort should be made to measure the compound or its metabolite(s) in blood and excreta. Such measurements permit use of information gained during the test in the solution of diagnostic and other practical problems. For example, if a person is suspected of being poisoned by a particular compound, it is of great interest to know whether the concentration of it in the blood is merely equal to or is significantly greater than that previously found in healthy volunteers or workers. By much the same token, it is valuable to learn as much as posssible about the metabolism of compounds in humans. Information on this matter may help to explain observed similarities or differences in the pharmacological or toxicological action of the compound in humans and animals (Brodie, 1964). In particular, such information may help in selection of the best species of experimental animal for further exploration of the compound. It often happens, not only in connection with metabolism, but also in connection with other parameters,

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that studies in humans will raise a question that can be explored best in animals before it is returned to humans a second time for further evaluation. No matter what special parameters are selected for study, every test in humans must include a medical history and general physical and laboratory examinations. These studies must be made before the chemical under investigation is administered, and they must be repeated at suitable intervals. This general surveillance should detect unexpected reactions, including sensitization that may be peculiar to humans. The surveillance should also detect illness that is etiologically unrelated to the chemical under study. Such basically unrelated illness may influence the reaction of the body to the compound under study, or, if not detected early, it may serve to confuse interpretation of the study.

59.4.3.2  Selection of Dosage The highest dosage selected for a test in volunteers should be one that is believed to be (a) harmless on the basis of animal studies and all other available information and (b) capable of influencing some measurable parameter. For compounds of low toxicity, measurement of excretion of the compound or its metabolites may be the only parameter influenced. For compounds of moderate or high toxicity, it may be desirable to use a graduated series of dosages beginning with a very low one, and gradually approaching the highest dosage that produced no significant injury in animals. Regardless of the toxicity, it is usually desirable to have more than one dosage, either in a single test or in succeeding tests, in order to determine whether responses are dosage related. Even if dosage response is not under investigation, it is necessary that a control group be included in every study. Its use may serve to detect laboratory error and other factors that might influence the results without having any relation to the agent under study.

59.4.3.3  Choice of Volunteers Selection of volunteers is determined by (a) the number of people required and (b) the duration of the study. People from any group including the general population may be recruited for short periods of testing, especially if the tests involve little inconvenience. In studies in which each individual need be tested only once, and then only briefly, one could eventually study large numbers of people enrolled from the general population. However, for studies requiring a substantial number of volunteers for an extended period of time, the only practical method is to recruit volunteers through employment or through appeal to some special interest of the candidates. Thus, depending on the nature of their interests, volunteers for most studies of safety may include (a) patients, (b) paid employees from the general

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population, groups of workers, or other groups, (c) laboratory staff, (d) medical or other students, (e) military personnel, including conscientious objectors, (f) persons with some occupational interests in the problem under study, and (g) prisoners. If patients wish to volunteer for a test having no relationship to relief of their own condition, they may do so provided they are capable of giving consent. In fact, there are a few studies that in humans can be carried out only in patients. An example would be measurement after death of the distribution of a chemical in all tissues of the body. Another example would be any study involving debility or a specific metabolic disorder as a necessary condition. In any study of patients as volunteers, the investigator should not be the personal physician for any volunteer, and a personal physician as well as the investigator should be able to discontinue the participation of one or more volunteers as required. Of course, each patient should be able to stop participating at will. Fully paid employees have been utilized in small numbers, usually for special purposes. Although perhaps not typical, Alexis St. Martin, who was studied by Beaumont (1833), is the most famous employee hired in order that he might be studied. This method of obtaining volunteers deserves more consideration. It offers a particularly good opportunity to choose the desired distribution of persons of different age and sex. A number of studies have been carried out using laboratory staff. This method has the advantage that the volunteers are conveniently available for repeated brief examinations that, under other circumstances, would require far more of their time and of the investigator’s time. Another advantage is that such volunteers, because of their technical backgrounds, are in an unusually good position to understand the nature of the test and its potential risks. This means that the investigator need not exaggerate the danger of the study in order to be sure it has not been underestimated. The chief disadvantage of using staff as volunteers is that it is difficult to avoid subtle coercion. Many staffs are not large enough to provide a numerically adequate group for study even if everyone participates. The result is social pressure to be one of the group. The pressure may increase as the study progresses and the volunteers develop an esprit de corps. By the same token, a volunteer cannot leave such a study without attracting special attention. There is a certain danger that, because of their social and economic relationship to the investigator, these volunteers may be more likely than some others to behave as they think they should. Obviously, this danger is greater in connection with subjective tests and should be essentially nonexistent in connection with chemical or physical measurements. All the disadvantages of studies using laboratory staff are accentuated in studies in which the investigator is also a subject.

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The comments in the preceding paragraph apply with equal force to medical or other students used as volunteers. This is especially true if the students come from the same school or department as the investigator. The disadvantages would be lessened if the volunteers were drawn from a large student body so that, although the participants become known to one another, they are not generally recognized as a group by the student body. The use in medical studies of military personnel or civilian personnel at military installations is limited to the armed forces and the problems peculiar to their mission. People may volunteer for tests because of an occupational interest in the problem under study. In this respect, they are like military volunteers because they know they may contribute directly to their own future welfare if they can help in the solution of a problem peculiar to their work. For example, crop duster pilots volunteered, during the winter season when they were not flying, for a study on how tetraethylpyrophosphate (TEPP) affects vision and how, when TEPP contacts the eyes, it interferes with the pilots’ ability to judge distance and thus to control their planes (Upholt et al., 1956). Prisoners offer many advantages as candidates for medical study. The population of a single prison is often large enough to permit an adequate number of volunteers of the desired range of age. In many instances, both men and women will be available through the same prison administration, though seldom at the same installation. If the institution is large enough, sufficient volunteers may be obtained without exerting any social pressure on them, and a volunteer can stop participating without drawing special attention from his fellow prisoners. With proper scheduling, the studies may be carried out with efficient use of the investigator’s time, and without interfering with the prisoners’ regular duties. The disadvantages of using prisoners include their erratic motivation and their often considerable variation in ability to understand the nature and risks of a study. A major disadvantage is that all doses or other treatments must be given under observation. By contrast, staff members not only may take their own doses but may do so during travel or under other inconvenient circumstances. Those without personal familiarity with tests in humans sometimes express doubts that volunteers act without some form of coercion. The doubt is probably most often expressed about prisoners, but may refer to medical students or any other group from which volunteers may be drawn. The most convincing argument against this sort of doubt is active experience with a properly conducted study. Recent evidence that prisoners may truly wish to serve is the fact that, in response to bureaucratic action, one group of them sued the federal government of the United States to retain the right to volunteer (Sun, 1981).

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59.4.3.4  Studies of Workers as Volunteers In the preceding section on volunteers, workers might be included in several categories, especially “paid employees,” although the pay in question is that received for medical study and not that for whatever duty makes the person a “worker.” The distinction between participation of an employee in a study as a volunteer or as a worker lies in the nature of the exposure. Regardless of his employment status, a person is a volunteer if he agrees to receive an exposure administered for the purpose of making a study. On the contrary, a worker is not a volunteer in the usual sense if he agrees to undergo medical study in relation to an exposure or some other factor that is a necessary condition of his regular employment. One might ask how participation of a worker in the kinds of tests under discussion differs from that ordinarily involved in good industrial medicine. The different is one of emphasis; there is no clear dividing line. Tests in humans emphasize our ignorance of how chemicals or other environmental factors act in humans; hence, these tests seek new information. They seek safety by trying to identify and define unexpected dangers. Industrial medicine emphasizes our wealth of knowledge of how chemicals or other environmental factors act in humans. Like any good medical practice, industrial medicine attempts to prevent – and, if necessary, cure – injury by taking all known dangers and signs of danger into account. Some opposition to tests in humans seems to be based on a desire of industry to limit the extent of tests, of whatever nature, that are required for official acceptance of a product. This opposition might be reduced if it were understood fully that many of the tests that need to be done regarding pharmacology and safety can be carried out in workers as an extension of good occupational medicine. It is a curious fact that, in studies of workers, no question is raised about consent for exposure although consent for diagnostic procedures should be obtained by the investigator. It is assumed that consent for exposure exists if a person is employed in the manufacture or formulation of a chemical, provided the management takes conventional measures to limit exposure and no illness occurs as a result of the exposure. This is true even in those instances in which the amount of chemical absorbed is equal to or greater than that adequate for studies in volunteers.

59.4.3.5  Protocol and Conduct of the Study (a)  Function and Content of the Protocol The plans for every study in humans should be put in writing. Before this is done, an informal agreement with the administration of any institution where the study is to be made ought to be reached. Such preliminary discussion may save a great deal of inconvenience by exposing requirements of the institution that, although of no scientific

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importance, may be of crucial practical significance. In general, the writing of a protocol forces the investigator to face not only the scientific but also the technical and administrative logistics of the study. Once the document is prepared, it may be used as a basis for approval of the project by (a) management or custodial authorities for administrative feasibility or (b) academic or comparable authorities for scientific merit and safety. The protocol should contain at least (a) a summary of the problem that justifies the study, (b) the specific purpose of the study, (c) experimental design, (d) safety of the study, (e) rewards to volunteers, and (f) a proposed contract or other form to document consent. The “problem” referred to in item (a) may be very broad. It is often ill defined, but never unimportant. An example would be the need to control the vector of urban yellow fever, dengue, and hemorrhagic fever. The fact that the vector, Aedes aegypti, is not yet fully controlled only adds to its importance as a medical problem. By contrast with the problem, the purpose of any study in humans should be clearly defined and is frequently very limited. Thus, study of a candidate insecticide such as difenphos to help to control A. aegypti might be limited to investigation of its metabolism and storage in normal people. The section of the protocol dealing with experimental design should state the number of people to be studied, their distribution into experimental groups, the rate and duration of dosage, and the observations to be made. The groups should be large enough to permit statistically valid conclusions, even if some of the participants drop out before the study is complete. The protocol should state clearly what measurements are to be completed before dosage or some other experimental situation is instituted. The protocol should also state what observations are to be made after dosage is discontinued. The duration of study need not be predetermined, but the schedule may be defined in terms of some pharmacological constant. For example, the study may be continued after dosage is stopped until the excretion of a metabolite of the chemical under study reaches the level found in the controls. The section on experimental design should give particular attention to aspects of the study that directly concern the volunteers. Thus, the protocol should state the kind and approximate number of samples that will be taken from each person. It is just as important to enumerate procedures that may involve inconvenience or interference with institutional routine as it is to define those that may involve pain, discomfort, or even risk. The author and his colleagues almost lost a project because the collection of the urine samples interfered with the employment of a few of the volunteers in a weaving mill, which was a part of the occupational training and therapy program of the prison. Fortunately, it was possible to change the schedule for collecting samples without decreasing their number or scientific

Hayes’ Handbook of Pesticide Toxicology

value, and the study was saved. Not all difficulties can be foreseen, but they can be reduced by careful planning. The section on safety is one of the most important parts of the protocol. This section should be brief and clear. The major points may be supported by published reports of animal experiments or by other appropriate documents presented as appendices. In any event, this section must define (a) the nature of any hazard, (b) the probability of danger, (c) the means available to detect or, if possible, predict injury, and (d) the possibility of specific treatment if injury should occur. The section on safety must include a statement that individual volunteers or the entire group will be withdrawn from active participation if circumstances require it. Of course, a volunteer whose active participation is stopped will continue to be observed and to receive therapy if needed. The protocol should specify the rewards (e.g., money and a certificate) the investigator is prepared to give the volunteers. It may also list other compensations that would be appropriate, although they are not under the control of the investigator. For example, if the work is done at an institution that is able to offer some other form of reward, then that, too, should be mentioned. Some prison systems routinely give a reduction in sentence for good behavior, including but not restricted to participation in medical research. The last section of the protocol should refer to the agreement or contract to be used in connection with the study. The author and his colleagues have used an elaborate contract in connection with studies in institutions. The contract includes (a) the date, (b) the name of the organization (or persons) making the study, (c) the name of the institution where the study is made, (d) the name of the applicant, (e) a description of the study, (f) the kind and approximate number of services the applicant will be expected to contribute and the approximate duration of service, (g) a statement by the applicant that the procedure, value, and danger of the study have been explained, and a statement of full awareness that there can be no guarantee against illness (in a study in which illness really is anticipated, its nature should be defined in the contract), (h) a statement by the applicant of knowingly and voluntarily accepting the risks and of agreement to cooperate in the work, (i) the amount of money and other rewards the applicant will receive in exchange for the services, provided participation is satisfactory, (j) the applicant’s statement on behalf of self, representatives, and heirs of release of the investigators and the organization they represent from all liability, including claims and suits at law or in equity for any injury, fatal or otherwise, that may result from participation in the investigation, (k) signature of the applicant, (l) signature of witnesses, (m) signature of the responsible investigator who accepts the applicant as a volunteer, and (n) signature of the officer who approves the contract for the institution where the work is done.

Chapter | 59  Studies in Humans

The contract always omits any statement that the applicant agrees to remain in the project. In some instances, the signatures mentioned above are followed by this statement signed by an officer of the institution. I attest that the purposes, procedures, and inherent risks of this study were fully explained to the applicant by the officer in charge of the investigation. I am convinced of the applicant’s complete understanding of the study and of his willingness to participate in it. I am confident that no duress of any kind was present in the proceedings leading to the signing of this contract.

(b)  Initiation of a Study The method of obtaining volunteers must vary with circumstances. If the study involves members of a laboratory staff, they will be generally familiar with the need for the study and the character of the risk, if any, as the preliminary work progresses in experimental animals. If the work is to be done with a group of people who have no preliminary introduction to it, it is best that a completely impersonal notice be issued and that the candidates consist only of those persons who show enough interest to seek further information. Such a notice should state briefly the purpose and nature of the study and invite anyone interested to attend a meeting at a specified time and place at which a further explanation will be offered. If the work is done in a prison the custodial staff should screen those who propose to attend the introductory meeting in order to exclude those who are unsuitable by reason of a short sentence, a reputation for making trouble, or some special duty that would conflict with participation in the study. At the meeting, the principal investigator (and others if necessary) should explain the study in detail. Emphasis should be placed on (a) the value of the work, (b) what each volunteer will be asked to do, and (c) the hazards, no matter how remote, that the volunteer may encounter. Of course, the rewards, the agreement, and other administrative arrangements should be explained. When questions from the candidates have been answered, the candidates are asked whether they wish to volunteer. A list is made of those who do, and they are instructed when and where to report for preliminary medical examination. The preliminary examination consists of a complete medical history and appropriate physical and laboratory examinations of each applicant. The examination also involves special tests, including the index test, dictated by the nature of the study. One or more parts of the preliminary examination may serve to exclude some applicants from the study, and all parts are used as a baseline for judging future findings in the others. Those candidates who are found completely satisfactory are asked to sign the agreement or contract, and they are not considered volunteers until they have done so. Although candidates for a study should be told its complete design, it should be understood that no volunteer is

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to be told the dosage group to which he or she belongs. A useful device is to give placebos to all dosage groups for about a week at the beginning of a study. Appropriate samples are taken during this period when each volunteer serves as his or her own control. Sometimes more complaints are received during the time placebos are given to everyone than during the remainder of the study. This is entirely understandable because the volunteers have been told about risks that the investigator felt morally obliged to mention, although the chance that they would materialize was considered really nonexistent. Furthermore, it is fair to say that volunteers think that their appetite or general health has improved about as often as they think they have suffered some injury. In any event, it is absolutely necessary to have one or more control groups throughout the study. Such a group permits a random distribution of complaints that have no real connection with the experimental conditions. It also permits immediate recognition of the cause if some unexpected finding is the result of laboratory error. Some measurements tend to vary even when made in the best laboratories. An example is the potentiometric measurement of blood cholinesterase. If it is found that the blood cholinesterase of all groups, including the control, drops sharply for a day or two and then returns to normal, the result may be safely ignored as being unrelated to the variable under study. (c)  Double-Blind Studies The double-blind technique requires that neither the subjects nor the investigators responsible for recording their reactions be aware of which volunteers receive different chemicals or different dosage levels (including placebo) of the same chemical. It follows that a third party will be responsible for assigning dosages but have no direct contact with the volunteers during the course of the study. The third party, usually a member of the team of investigators, may take major responsibility for design of the study and interpretation of the results, but not for collection of the data. The great value of the double-blind technique in many situations should not lead to a routine demand for it in all studies. Need for the technique diminishes as the degree of difference to be measured increases. Furthermore, the need is inversely proportional to the objectivity and repeatability of the measurement required. Finally, the double-blind technique is contraindicated in connection with early studies of really new materials. Under these circumstances, the investigator should have every opportunity to discontinue or otherwise modify the study in response to the first hint of injury. Contrary to the opinion of some authors, use of the single- or the double-blind technique has no necessary bearing on the matter of consent. Candidates for studies by the author and his colleagues were always told what dosage levels were planned, including the fact that some

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volunteers would receive a preparation containing none of the chemical under study. The only information withheld was who got which dosage. The candidates were told not to volunteer unless they were willing to receive the highest dosage. (d)  Closing Ceremony It is almost always wise to have a closing ceremony to recognize the contribution of the volunteers. The main purpose is to thank them and to emphasize the contribution of the volunteers toward solution of a scientific problem.

59.4.3.6  Protection of the Volunteer Protection of the volunteer, to summarize what has already been said in other contexts, includes (a) guarantee that no study is made without valid consent, that is, continuing consent freely given by a person legally capable of giving it (the recognized ability of the volunteer to withdraw from the study at any time, even without reason, is a part of this concept), (b) qualification of the investigator, (c) freedom of the investigator to stop the participation of volunteers if it is possible that their continuation in the project might injure them, and (d) care in the event of injury. Tests in humans should not be done in situations that do not permit medical care of a volunteer in the event of illness. For some large medical institutions, this presents no problem. For example, provisions for care of whatever duration necessary is made in the Army Regulations cited earlier. For smaller institutions, insurance may be the best method to prepare for the unlikely eventuality of injury. M. R. Zavon (personal communication to W. J. Hayes, Jr., 1964) reported that, in a study of the irritant effect of air pollution on the eyes, insurance for $3 million was purchased for $278.90.

59.4.3.7  Protection of the Investigator Protection of the investigator should be built into the design of every test in humans. Of course, the most important protection of the investigator consists of demonstrable evidence that he or she has done everything possible to protect the volunteers. (a)  Proof of Consent As discussed earlier, only the matter of consent was brought into question in the case of Hyman v. Jewish Hospital (15 NY 2d 317). Proof of consent involves not only evidence of the fact but also evidence that the volunteer was legally capable of giving consent. The fact of consent is best established by a written document as discussed under Section 59.4.3.5. While written consent will not protect the investigator against liability for negligence, it will constitute a defense against claims for injuries resulting from foreseeable and assumed risks (Monthly Summary

Hayes’ Handbook of Pesticide Toxicology

from the Office of General Counsel, U.S. Department of Health, Education, and Welfare, December, 1959). (b)  Proof of Qualification Any investigator or team of investigators undertaking tests in humans should be fully qualified not only from a general medical standpoint, but also from the special standpoint of the problem under study. Evidence of qualification may include consultantships, membership in learned societies, and papers published in the field. Proof of qualification for a particular study is strengthened by a signed permit from the investigator’s institution and/or another agency involved. Such approval would, of course, be conditioned on other relevant factors as well as the qualifications of the investigator. (c)  Proof of Value of the Research In many instances the value of research in humans is evident and requires no justification. Even in such instances, the reason should be outlined clearly in the protocol. Again, official approval offers additional evidence that the work is considered valuable. Because the word “experimentation” carries the legal implication of malpractice, it is only common sense to avoid its use in connection with proper medical investigation. (d)  Insurance Insurance is a protection not only for the volunteer but also for the investigator (Beecher, 1969). Ballard (1964) has discussed the possibility of legal and financial support from the manufacturer of a compound in the event that an authorized investigator of the compound is sued. The rate of insurance would be influenced much more by the volume of sales than by the small risk involved. A survey of 39,216 subjects participating in therapeutic research indicated trivial injuries (nausea, vomiting, headache, hematoma, etc.) among 8.3% of them, temporarily disabling injuries in 2.4%, permanently disabling injuries in 0.1%, and fatal outcome in 0.1%. Among 93,399 subjects participating in nontherapeutic research (including some diagnostic or other studies of the preexisting illness of some patients), there were trivial injuries in 0.7%, temporarily disabling injuries in 0.1%, permanently disabling injuries in 0.1%, and no fatalities; the one case of permanent disability was a stroke, not clearly related to the research, that occurred 3 days after the investigative procedure ended (Cardon et al., 1976). (e)  Tort Claims Act For many years there was a need for tort claims acts to protect government-employed physicians and other medical personnel acting within the scope of their employment, including the conduct of medical investigations (Hayes, 1965, 1968). Such a law for Defense of Certain

Chapter | 59  Studies in Humans

Malpractice and Negligence Suits now has been approved in the United States. It was a portion of the Emergency Health Personnel Act of 1970 and constituted Section 223 of Public Law 9 1-623, approved on December 31, 1970. The present wording is recorded in 22 USC 2702. The act provided protection from suit or damage for personal injury, including death, resulting from the performance of medical, surgical, dental, or related functions, including the conduct of clinical studies or investigation, by any commissioned officer or employee of the Public Health Service while acting within the scope of his office or employment. The law has been extended and now provides protection of all federal medical personnel acting within the scope of their employment even if on assignment to states, foreign countries, or nonprofit institutions. Although the law and foreseeable extensions of it apply directly only to government employees, it may be expected to influence legal action at all levels. The specific inclusion of the “conduct of clinical studies or investigations” in the act is a substantial evidence of confidence in medical research in general.

59.4.4  Motivation of Volunteers Different people participate in research for different reasons or different combinations of reasons, including (a) tangible rewards, (b) a desire for new experience, (c) occupational or other specialized interest in a specific project, (d) interest in science generally, and (e) a desire to contribute to society. Tangible rewards may include money and (in the case of prisoners) reduction of prison term. Reduction of sentence may be guaranteed or potential; that is, it may be prorated according to the duration of each volunteer’s participation in the study or it may consist of the hope that participation will influence parole or the dropping of detainers by another court. Although tangible rewards may be critical for participation in some instances, the importance of this factor is easily overestimated. The author and his colleagues have carried out a number of studies successfully even though only token payments were affordable. We have recruited prisoner volunteers without difficulty when no reduction in sentence could be offered. A general interest in science is probably the most important overall reason leading people to become volunteers in medical studies. Many people want to help solve any problem they understand. If the problem involves their occupation, there is every reason they will understand its importance, and they may even have some technical insight into it. Of course, prisoners share with people in general an interest in science, a need for money, and a desire for new experience. In fact, the desire for new and varied experience may be especially strong in people who are confined. However, the main feature that distinguishes prisoners from others in this connection is an unrecognized urge

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to do something useful. Either on a conscious or subconscious level, many prisoners wish to compensate for records indicating little positive contribution to society and, consequently, little positive recognition from society. Few prisoners will admit this urge to contribute except in time of war, when any worthwhile act can be attributed to patriotism. A partial exception was revealed by one prisoner who told the author that participation in a project had led to his reinstatement in his family. His father was pleased that he had finally done something commendable. Continuation in a project requires an element of steadfastness that is not necessary for initial recruitment. This steadfastness seems to be characteristic of the individual; it has little or no relation to the nature of the study or to whether discomfort is or is not involved. Dropping out must be permitted, but in the author’s experience it usually occurs very early in a project, frequently before anything has happened. Once a project has settled down to a routine, it is most unusual for a volunteer to drop out. In studies such as those on antimalarial drugs, in which production of the disease was a necessary part of the work, the volunteers accepted their illness with true heroism.

59.4.5  Studies of Pesticides in Volunteers Compounds that have been studied in volunteers are listed in Table 59.2. This list excludes studies in workers, including workers involved in community studies (the so-called “village trials”) conducted by World Health Organization teams. The list includes some preliminary studies of pesticides intended as drugs but excludes therapeutic evaluations of them in patients. Each published report of a study of a compound in volunteers constituted a unit for tabulation. If a paper reported study of two compounds, it appears twice in Table 59.2; however, most papers reported only a single study. The 312 reports tabulated were contained in 264 papers. In Table 59.2, each report is assigned to the single kind of study most emphasized in it. “Metabolism” includes absorption, distribution, and excretion as well as biotransformation. “Safety” is divided into general, pharmacology (including measurement of cholinesterase activity), and “sensation,” involving mainly irritation and sensitization but also a few studies on odor and taste. Of course, the studies of metabolism may have important implications for safety also. Some persons may be astonished that at least 102 pesticides or related compounds have been studied in volunteers. (Compounds considered as related include one synergist, one repellant, and one toxic by-product of chlorinated phenols.) The nationality of origin of the 312 studies of pesticides was determined according to the nationality of their authors, which in most instances, corresponded with the country where the work was done. However, there were six reports

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Table 59.2  Number of Studies of Pesticides and Related Compounds in Volunteers Class and compound

Study

Class and compound

Safetya

Metabolism G

P

Metabolism

S

Inorganic compounds

Study Safetya G

P

Organic phosphorus compounds (continued)

Alkyl mercury compounds

1

Diazinon

1

Arsenic

4

Dichlorvos

8

Barium carbonate

1

Dimefox

1

Boric acid

2

Dimethoate

2

Elemental mercury

4

Dioxathion

1

Elemental sulfur

1

EPN

3

Lead

3

Fenitrothion

1

Lead arsenate

1

Malathion

3

Mercuric chloride

1

Methidathion

2

Thallium sulfate

1

Mevinphos

4

1

Botanicals 2

Bacillus thuringiensis

2

4

Monocrotophos

1

2

Parathion

6

6

Parathion-methyl

Barthrin

Phenthoate

1

Phosphamidon

2

Fenproponate

1

Pirimiphos-methyl

1

Fenvalerate

2

Schradan

2

Flucythrinate

1

Temephos

Cypermethrin

2

Ivermectin

1

Nicotine

1

1

2

1

TEPP 4

Permethrin

3

12

1

3

Trichlorfon

2

6

3

Phenothrin

1

Carbamates

Pyrethrum

3

Aldicarb

Rotenone

2

Bendiocarb

1

1

Carbaryl

4

4

Carbofuran

1

1

Salmonella

Solvents and fumigants Carbon disulfide

S

1

formetanate 3

propoxur

1 3

1

Chapter | 59  Studies in Humans

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Table 59.2  (Continued) Class and compound

Study Metabolism

Carbon tetrachloride

3

Class and compound

Safetya G

P

Metabolism

S

1

Study Safetya G

P

Thiodicarb

Chloropicrin

S 1

2

Dichlorodifluoromethane

2

Dichloromethane

4

1

1

Phenols DNOC

1,3-dichloropropene

2

1

Dinocap

1

Epoxyethane

1

Pentachlorophenol

2

Isobornyl thiocyanate

1

TCDD

1

Tetrachloroethylene

4

2

1, 1, 1-trichloroethane

5

4

Trichloroethylene

6

1

Xylene

2

5

1

4

1 Synthetic rodenticides 4

Cholecalciferol

1

Norbormide Warfarin

1

1

6

Chlorinated hydrocarbons BHC and lindane

4

DDT

7

Dieldrin

3

Herbicides 13

2

1

1

Methoxychlor

2

Toxaphene

1

Cycloate 2,4-D

1

1 4

1

Dicamba Diquat

1 1

Glyphosate Organic phosphorus compounds Azinphos-ethyl Azinphos-methyl

1

Carbophenothion Chlorfenvinphos

1

Chlorpyrifos

1

Chlorpyrifos-methyl Demeton DFP

1

1

Paraquat

1

1

Picloram

1

4

Silvex

1

1

2,4,5-T

2

2

Fungicides

1

Captafol

1

7

Captan

2

1

Dichloran

1

1 (Continued )

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Table 59.2  (Continued) Class and compound

Study Metabolism

Class and compound

Safetya G

P

Diphenyl

Safetya G

P

S

Miscellaneous compounds (continued) 1

Niclosamide

Quintozene Thiabendazole

Metabolism

S

Fungicides (continued)

Study

1

2

Piperonyl butoxide

2 1

1

Zineb

1

Total (102 compounds)

120

61

93

38

Miscellaneous compounds Amitraz Deet

1 5

2

a

G, general safety; P, pharmacological study; S, sensation (mainly irritation and/or sensitization but a few related to odor or taste).

of studies carried out at navy bases or civilian hospitals in the tropics by investigators from Belgium, the United Kingdom, or the United States. The origin of three studies was undetermined. The majority of studies (202) were by authors from the United States; the next largest number (30) were from the United Kingdom. Studies by authors from individual countries were distributed as follows: North America, 5; South America, 1; Europe, 63; Africa, 4; and Asia, 7. The failure of the sum of origins (known or undetermined) to correspond with the total number of reports reflects the international origin of three of the studies. A study of pesticides in volunteers occurred as early as 1880 and again in 1893, but the next was in 1922 (Figure 59.1). Such studies were unusual until World War II. Many of the wartime studies were directly related to the protection of troops from vector-borne diseases. Some of the reports were published during the war, and others were published soon after hostilities ended. Such studies then decreased gradually; none were recorded in 1955. Beginning in 1956 there was a striking increase in studies in volunteers until 1969, when there were 14 papers reporting on 18 compounds. For a time a high level of research was maintained; for the 5-year period 1969–1973, the average number of studies per year was 13. The number of studies then declined and later evened off so that the average number per year for the 5-year period 1983–1987 was 5.4.

59.4.6  Conclusion The limited ability of animal tests to predict human reactions to chemicals is recognized generally. Tests in humans represent a logical step toward full-scale use of a chemical. The fact that even hypersensitivity is basically dosage

related emphasizes the importance of a continuing surveillance of workers or other persons whose exposure is intensive and prolonged in order to detect idiosyncrasy as well as overt toxicity at the earliest possible moment. Many countries accept the necessity for orderly testing of the effectiveness of new drugs in humans. Such testing of drugs for effectiveness implies, but does not require, that the safety of therapeutic doses will be observed critically and that some pharmacological studies of the drug will be made in humans. Furthermore, testing for effectiveness does not necessarily mean that possible adverse reactions will be looked for systematically after a drug is released for prescription use. Thus, studies of the safety of drugs in humans often are not required by law or custom in exactly the same way that tests of effectiveness are required. There is even less recognition of the need to test in humans the safety of other chemicals to which people will be exposed. In fact, little thought seems to have been given to the question of whether it is ethical to release a new chemical for general use before it has been tested under controlled conditions in a limited population. However, there is some evidence – including circumstances surrounding the case of Hyman v. Jewish Hospital – suggesting that not only investigators but also many other educated people consider that, as long as the subjects of research are uninjured, it is ethically more important to gather knowledge leading to human welfare than to observe legalistic limitations of informed consent. If it is decided that general release of a new chemical under conditions that defy orderly study is unwise – or, in fact, if studies of therapeutic effect are to be done scientifically using proper controls – then one must abandon the notion that tests in humans can be justified only if they

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20

Number

15

10

5

0 1920

1930

1940

1950

1960

1970

1980

Year Figure 59.1  Year of publication of studies of pesticides in volunteers. , published papers; , compounds reported in papers reporting more than one compound. Modified from Hayes (1983).

are of possible direct benefit to each person participating in a test. Countries apparently vary widely in the tenacity with which they support this notion. Certainly, tests of great value to society have been carried out without injury, but also without any medical benefit to the volunteers involved. However, only in special situations, especially in the military, are such studies clearly permitted by law. Great progress was made by the recent passage of a law to protect any commissioned officer or employee of the Public Health Service against suit for damage resulting from action within the scope of his or her office or employment, including those actions involving clinical studies or investigations. Further progress could be made by passage of additional tort claims acts to protect all government employees with medical duties, including duties of medical research. This would serve not only to protect the employees but also to establish the legitimacy of their proper activities. Such laws might serve as devices to define standards of conduct for investigations. However, the present law makes no such attempt. This may be just as well. In the last analysis, the proper conduct of tests in humans depends on the good sense, integrity, and freedom of the participating scientists, rather than on an elaborate code of ethics.

chemicals, including pesticides. It can be argued that the only really important dosage is one at the tissue level and that it is equally easy to measure this level regardless of the kind of exposure. However, it is also important to know the dosage of a compound encountered by the entire body and to know the relative importance of different routes by which it may be absorbed. This section describes methods for (a) the direct measurement of exposure, (b) certain special applications of indirect measurement aimed at evaluating the separate contributions from different routes of exposure, and (c) measurement of absorbed dose. Practical methods of indirect measurement of exposure resulting from occupation or accident include direct measurement of storage or urinary excretion. Another method, measurement of respiratory excretion by breath analysis, although relatively new, has been brought to a high degree of perfection for many solvents and fumigants. It deserves far wider use than it has received. However, respiratory excretion can be used in studies of volunteers or as a part of good occupational medicine (Stewart, 1974).

59.5  Measurement of exposure and dose under practical conditions

Respiratory exposure cannot be separated completely from oral exposure because some inhaled material is likely to be retained on the mucous membranes of the upper respiratory tract. This material may be absorbed directly from these membranes or swallowed so that it becomes available for absorption from the gastrointestinal tract. The depths which inhaled material will reach in the respiratory tract are determined largely by particle size.

Because of the crucial importance of dosage in all toxicological matters, it is imperative that dosage be known wherever possible. This measurement is relatively easy in connection with therapeutic drugs but difficult in connection with occupational or accidental exposure to industrial

59.5.1  Measurement of Respiratory Exposure

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59.5.1.1  Estimation of Respiratory Exposure from Air Concentration It is common practice to estimate the respiratory exposure of workers in factories and mines by measuring the concentration of toxicant in samples of ambient air, and by assuming that the samples are representative of the air inhaled. This assumption is generally valid when the concentration of toxicant in the air is relatively constant and when the toxicant itself is a gas, vapor, or extremely fine particulate. The assumption that ambient air is representative of inhaled air may be misleading when the concentration of toxicant in it varies greatly from time to time or when at least a portion of the toxicant is present in the form of relatively large particles. If the toxicant enters the air in the form of clouds that quickly drift away, it may be difficult to obtain a representative sample. If the cloud lasts only a very short time, the worker may be able to hold his or her breath or turn his or her head so that little of the material is inhaled. If a cloud of whatever duration is made up mostly of relatively large particles, then only a small portion of the toxicant present can reach the lung even if the large particles are inhaled. In spite of these necessary limitations, measurement of air concentration is an important way of estimating respiratory exposure. No matter what device is used in collecting air samples for analysis, an effort should be made to obtain samples with the same spectrum of particle size as is picked up by the nostrils in breathing. One approach is to imitate the aerodynamics of human respiration. For example, it may be possible to make the position and size of the collecting orifice similar to the nose and to make the rate of air flow similar to that in human respiration. In all sampling, the collecting orifice should be as near the breathing zone of workers as is practically possible. If very large particle sizes are involved, it is important that the orifice be inverted so that large particles will not fall in simply mechanically. After all, some spray droplets approach the size of fine rain droplets, which are too heavy to be diverted from their free fall and be caught up in the respiratory stream. Regardless of what collecting equipment is chosen, it is wise to use all glass fittings. Traces of material obtained from rubber or plastic produce false-positive results for many pesticides when measured by gas–liquid chromatography. In selecting a liquid to use in impingers or fritted glass absorbers, it is necessary to consider vapor pressure as well as solubilizing action. Ether and acetone generally are not satisfactory, even for collecting compounds highly soluble in them, because these solvents evaporate too rapidly. The material to be sampled may codistill with the solvent. On the other hand, use of a too-viscous solvent will impede the flow of air through the sampling device. Miles (1965) has shown the value of ethylene glycol as a medium for collecting parathion. Although this particular solvent might

not be ideal for other materials, the principle of seeking maximal dissolving power and minimal evaporation is always valid. Another approach is to use some kind of a filter instead of a solvent. It might appear that filters would act mechanically and trap only particles larger than the pores of the filter, thus failing to capture smaller particles, gases, and vapors. Actually, the situation is far more complex. Extremely fine solid particles are more difficult to trap than larger particles on the one hand or vapors on the other. In some situations, a high proportion of small particles will be carried directly through an impinger filled with a solvent in which the material composing the particles is soluble. On the contrary, vapors and gases may be adsorbed on filters or dissolved in solvent. Surface area is the most important single factor in determining whether adsorption will occur. Thus, filters made of glass wool usually adsorb less gas or vapor than columns packed with alumina. Of course, the specific properties of the adsorbent are also important; therefore, the efficiency of a filter or column packing should be tested before it is selected for a particular purpose. After the concentration of the toxicant in the air has been measured, the respiratory exposure of workers or others can be calculated by using an assumed minute volume. The average values for lung ventilation in adults at different degrees of activity are given in Table 59.3. Minute volume varies not only with activity but also with body weight, as shown by the fundamental studies of humans and laboratory animals by Guyton (1947a,b). Thus an even more accurate calculation can be made if lung ventilation is actually measured under the conditions of work being studied.

Table 59.3  Average Lung Ventilation in Young Athletesa Sex

Condition

Rate (liters/min)

Male

Rest

7.4

(5.8–10.3)

Light work (500 kg-m/min)

28.6

(27.3–30.9)

Heavy work (800 kg-m/min)

42.9

(39.3–45.2)

Rest

4.5

(4.0–5.1)

Light work (300 kg-m/min)

16.3

(15.9–16.8)

Heavy work (600 kg-m/min)

24.5

(17.3–31.8)

Female

a

Data from Taylor (1941).

Chapter | 59  Studies in Humans

(a)  Collection of Particles by Impaction on Screens It may be desirable to learn something, either qualitatively or quantitatively, about the dustiness of an environment in situations in which ordinary air sampling equipment is not suitable, either because of a lack of power to run sampling equipment or because the air flow associated with such equipment would introduce an important undesirable bias. Collecting screens have been used to meet problems of this sort. Apparently, the first use of screens in this way and certainly the most thorough study of their properties was that of Blifford et al. (1956). Their initial studies were with metal screens, which they compared with “an efficient filter apparatus” and with standard gummed-paper fallout sheets. They concluded that the screens were about 1% efficient in the absence of rain and that ordinary cheesecloth screens lost less material during rain. Their screens were mounted on rotating vanes to direct them toward the wind. However, no such mounting is required if screens are used to sample particulate material outdoors where the wind is always from the same direction or indoors where dust is carried by imperceptible random air currents. The lack of quantitation is a disadvantage, but it may be outweighed by other features of the system. The use of screens is a valid way to compare the dustiness of indoor situations or to compare the composition of dust in different situations. Fabric used for screens should offer little restriction to the passage of air and (following suitable cleaning) the fabric itself should not produce peaks on analysis. Nylon chiffon has been found suitable. The amount of dust retained may be increased by pretreating the clean fabric with 10% ethylene glycol in acetone (Tessari and Spencer, 1971).

59.5.1.2  Estimation of Respiratory Exposure from Trapping Toxicant in Inhaled Air A different approach for estimating respiratory exposure avoids some of the complications just discussed by relating the sampling directly to the worker or other subject.

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The method, which apparently was introduced by Batchelor and Walker (1954), involves determination of the amount of toxicant trapped by the filter system of a respirator worn by the worker during an accurately timed period of exposure. The nature of the apparatus is shown in Figure 59.2. If a properly fitted face piece is used, all inhaled air must pass through the filter system; thus, the volume of inhaled air does not have to be measured or estimated. The equipment is so designed that the air enters it through holes similar to the nostrils in placement and size. This may be achieved by firmly taping a modified funnel to the retaining ring of the respirator. The stem of the funnel is shortened and plugged, and two holes 12 mm in diameter and 6 mm apart are drilled midway between the base and the apex of the funnel. In use, the holes of the funnel are directed downward. The special funnel prevents the direct impingement of large particles on the filter and, by controlling the velocity of the air, also regulates the particle size of the material having access to the filter. The same authors reported studies showing that, under some circumstances, omission of a special funnel from the apparatus could lead to a value about 30% higher than the true value because so much material impinged directly on the filter pad. They also showed that measurement made with the special funnel in place gave results in satisfactory agreement with estimates made in the same situation by properly measuring the concentration of the toxicant in the ambient air. The fully developed equipment described by Durham and Wolfe (1962) has a filter consisting of 32 layers of surgical gauze stapled to an a-cellulose respirator pad. It was found that such a multilayer filter was efficient for trapping both sprays and dust. Tests indicate the filter pads could be relied upon to be about 90% efficient in removing parathion from air under the conditions ordinarily encountered in the field. Before filter pads are used, they should be pre-extracted with a suitable solvent to remove any material that might

Figure 59.2  Expanded view of a single-unit respirator equipped with a special pad to collect solid or liquid spray aerosols from inhaled air. On the front of the respirator is taped a plastic funnel with its stem plugged and with two openings 12 mm in diameter to admit air at the same velocity that the air enters the nostrils. The complete assembly as worn is shown at the lower left of the illustration. (From Durham and Wolfe, 1962, by courtesy of the World Health Organization.)

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later interfere with chemical analysis. To recover the toxicant after exposure, the entire filter pad is extracted in a Soxhlet apparatus. The amount of toxicant removed from the pad plus that rinsed from the inside of the funnel is considered the amount to which the worker would have been exposed by the respiratory route during the selected exposure period. The amount of toxicant received per hour, per working day, or per operational cycle can then be calculated.

59.5.2  Measurement of Dermal Exposure The classical problems of industrial toxicology involve respiratory absorption of heavy metals, toxic gases, and silica. All these materials are absorbed so much more efficiently by the respiratory tract than by the skin that dermal exposure to them may be ignored under practical conditions. Some of them such as silica have no systemic dermal toxicity regardless of the intensity of exposure. These classical problems of industrial toxicology remain of tremendous importance, even today. However, there is a growing list of industrial chemicals that may be absorbed from the skin efficiently enough that exposure to them by this route is dangerous under practical conditions of occupational exposure. Many of the pesticides fall into this class, and it is mainly in connection with pesticides that methods for measuring dermal exposure have been developed. In general, dermal absorption is far less efficient than respiratory absorption. Compounds vary greatly in the ease with which they are absorbed by the skin. In spite of these facts, dermal absorption is of crucial importance in some situations. It may happen that there is practically no opportunity for respiratory exposure in a situation that permits dangerous dermal exposure. More often, the ratio of respiratory to dermal exposure is such that the dosage resulting from the absorption of all inhaled material is less than the dosage from absorption of even a small percentage of the material contaminating the skin.

59.5.2.1  Estimation of Dermal Exposure from Air Concentration Edson (1956) proposed a table for conversion of air concentration values to surface contamination rates at various wind velocities. The table was based on the assumption that the surface would retain all spray particles approaching it at right angles. The table was used by Edson to show that, for various aerodynamic reasons, the surface contamination he measured was, in the absence of splashing, much less than would be predicted on the basis of the assumptions used. Even if more realistic assumptions were made, the prediction of dermal exposure from air concentration would be difficult, because the many different surfaces and

Hayes’ Handbook of Pesticide Toxicology

movements involved would create an extremely complicated aerodynamic situation. Because of the complexities involved, it seems impractical to use measurements of air concentration in estimating dermal exposure.

59.5.2.2  Estimation of Dermal Exposure from Absorbent Samplers Dermal exposure may be estimated directly by measuring the amount of toxicant trapped by absorbent pads or absorbent clothing worn by the operator in the course of work. These samplers can be attached to the hands, face, and other parts of the body that are normally unclothed. The samplers may be placed above and below clothing in order to estimate the degree of protection such clothing offers. No matter for what purpose they are used, the absorbent samplers are exposed for a carefully measured interval, usually one or more complete cycles of work. For example, measurements may be made from the time a sprayer begins to fill the spray tank with pesticide until that tankful of material has been completely used (whether for measuring direct exposure or the protection offered by clothing, gloves, or boots), as stated by Durham and Wolfe (1962). Briefly, it has been found practical to use pads made of acellulose for measuring sprays and pads backed by filter paper and made from 32 thicknesses of surgical gauze for measuring dust. Whatever material is chosen must be preextracted, using the same solvent to be used later during extraction of samples for analysis. It has not been found possible to remove inks and oils completely from samplers, so pads and gauze marked or treated with these materials should not be used. The exact size of pads is not important, but they should be uniform in size – small enough to put conveniently on the parts of the body to be studied and large enough to provide a margin by which they may be attached without interfering with a central portion later used for analysis. Square pads of a-cellulose (10.2 cm on each edge) have been found convenient. Gauze pads are backed by a single sheet of white filter paper and bound with pressuresensitive tape on all four edges in order to prevent raveling. When bound in this way, the gauze pad is slightly larger than the a-cellulose pad but not large enough to interfere with use. Laboratory tests have shown that gauze pads retain approximately 90% of dust applied to them, even though they are held in an inverted position and shaken in a mechanical shaker after application of the dust (Durham and Wolfe, 1962). Of course, in practice, all pads are handled with care to avoid loss of toxicant on the one hand and unintentional contamination on the other. It has been found practical to place exposed pads between the folds of a piece of quantitative white filter paper, which is then placed in a waxed paper sandwich bag for transport to the laboratory.

Chapter | 59  Studies in Humans

In the laboratory, each exposure pad is removed from the fold of filter paper inside the waxed paper bag. The margins of each pad are cut away with a paper trimmer, leaving a central area 5 cm on each edge. The 25-cm2 portion of the pad is then extracted in a Soxhlet apparatus. The total amount of toxicant finally measured in the extract is a measure of the exposure of the 25-cm2 area under the condition of the test. A shortcoming of the pad technique is the necessity of assuming that the area covered by the pad is representative of the entire body part being studied. This difficulty may be overcome by using knit white cotton garments that cover the entire body part during exposure. Gloves have been used for the hands, short-sleeved undershirts for the upper part of the body, socks for the feet, and the tops of socks for both ankles and wrists. The short-sleeved undershirts are of particular use for measuring the amount of toxicant that penetrates outer clothing, whether this be ordinary work clothing or special protective clothing. In a similar way, socks can be used to measure the amount of toxicant that penetrates shoes, and cotton gloves may be used to estimate the amount of exposure resulting from misuse of rubber gloves. Rubber gloves may be more permeable than supposed, or they may become contaminated on the inside through improper use. Occasionally, a garment may have to be cut off the subject in order to avoid its contamination by some part of the body over which it would normally be slipped in the process of removal. This is nearly always the case with the short-sleeved undershirt. However, the upper portion of a sock that has been used to measure exposure of the lower arm and wrist may be removed by slipping it over a plastic bag used to cover the hand. Uncut or resewn garments can be laundered and reused after they have been extracted to remove the toxicant under study (Durham and Wolfe, 1962). The ultimate extension of the use of garments for sampling exposure involves the use of coveralls. The protocol for measuring exposure issued by the World Health Organization (1982) contains a section concerning (a) measurement of total body exposure through the separate analysis of different parts of coveralls that cover normally bare as well as normally clothed parts of the body and (b) use of pads (or undergarments) to measure penetration of the coveralls by pesticides.

59.5.2.3  Estimation of Dermal Exposure from Washing It must be recognized that the ability of the skin to collect and hold a particular formulation of a toxicant may not be identical to that of an absorbent sampler chosen to study exposure. Some of the water-wettable powder that impinges on the skin may later dry completely and fall away. The degree of loss may vary depending

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on how wet the skin is with sweat. Thus, the efficiency of the skin as a collecting surface may not be constant. One approach to this problem is to use the skin itself as the collecting surface and to remove the toxicant from it by means of some kind of washing following a predetermined period of exposure. Although perfect in theory, this method, too, has some practical difficulties. For example, there is the difficulty of determining that the skin is completely clean before exposure starts, and the difficulty of knowing that all of the toxicant that impinged upon the skin during exposure is removed in the washing process. The skin is not inactive. In addition to possible loss through evaporation, the toxicant may be absorbed by the skin. Fredriksson (1961) showed that, whereas 80–92% of radioactive parathion applied to the skin of volunteers could be removed by ordinary washing 30 min after application, only 50–70% could be removed if the washing was delayed 300 min after application. In spite of recognized limitations, the method of washing has an important place in measuring the dermal exposure of occupationally exposed persons. There is a certain tendency for the two kinds of error mentioned above to cancel one another out. During relatively brief cycles of exposure, dermal absorption, though real, is minimal. In other words, because absorption is slow, the material that cannot be removed from the skin in preparation for measuring exposure may be almost the same as the material that cannot be removed after the exposure is complete. Thus, the quantity washed from the skin following exposure is a reasonable measure of the dosage received during the interval under study. Any part of the body can be washed with swabs even if the area chosen for study is small, large, or irregular. Durham and Wolfe (1962) recommended that each swab be made by placing two eight-ply, approximately 10  10 cm, pre-extracted surgical gauze sponges together and folding them twice to form an approximately 5  5 cm square that is stapled near the folded edge. The swabs to be used for a single area of skin are placed in a jar and saturated with 95% ethyl alcohol. The lid of the jar is lined with aluminum foil to prevent any possible extraction of the original liner. In the field, the sealed jar is opened and each swab is grasped at the stapled edge with ring forceps. The excess alcohol is squeezed off against the inside of the jar, and the area of skin to be studied is cleaned by rubbing the cloth back and forth over the surface with light pressure. After all the swabs have been used and placed back in the jar, it is sealed and returned to the laboratory for analysis. Durham and Wolfe (1962) found that one swab was insufficient, but four swabs were adequate to remove about 90% of parathion from an area of skin the size of the back of a human hand. It is difficult to clean between the fingers and around the fingernails with swabs. Durham and Wolfe (1962)

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described the use of plastic bags that they had introduced somewhat earlier for washing the hand and wrist or foot and ankle. In use, the hand is thrust into one of the bags containing about 200 ml of 95% ethyl alcohol or some other harmless solvent. While the open end of the bag is held tightly around the arm or wrist to prevent leakage, the thumb and fingers are rubbed briskly against one another and against the palm of the hand. Then the hand and bag are shaken vigorously about 50 times. The bags are carried to the place of sampling and back again to the laboratory in 0.5-l wide-mouthed canning jars. Both before and after use, the top of the bag is twisted to form a tight seal that is held with a clamp. Two bag rinses were found to remove an average of 96% of parathion from one hand when sampling was carried out soon after exposure, and two rinses were adopted as standard for this purpose. However, four rinses were found necessary to remove 90% of the recoverable toxicant one or more days after exposure. Durham and Wolfe (1962) found no significant difference in the efficiency of a-cellulose pads and swabbing for measuring insecticides on the dorsal surface of the forearm. However, bag rinses were approximately twice as effective as swabs in removing insecticides from the hands.

59.5.2.4  Conventions for Measuring Dermal Exposure Careful workers applying highly dangerous pesticides in a temperate climate will use protective clothing in such a way that their dermal exposure is slight. By contrast, some workers in tropical countries may wear little clothing regardless of the nature of the material they are applying. Their dermal exposure may be very great. Because of the wide variation in the kind and amount of clothing worn by workers, it is necessary to establish some standard for comparison so that measurements made under different circumstances may be evaluated. Thus, it has become conventional to calculate dermal exposure values on the assumption that the exposed person wore long pants, a short-sleeved open-neck shirt, shoes, and socks but no gloves or hat, and that this clothing gave complete protection of the area covered. This convention corresponds to approximately the smallest amount of protection used by workers in the developed countries. According to convention, the surfaces of the face, the back of the neck, the “V” of the chest, the forearms, and the hands are considered to be unclothed. The areas of these and other parts of the body are shown in Table 59.4. By use of published measurement of dermal exposure under standard conditions and the values for the surface area of different parts of the body shown in Table 59.4, one may get a rough idea of the added protection to be obtained from wearing protective clothing or the added danger that would follow the use of less than standard clothing.

Table 59.4  Surface Area of Portions of the Body Frequently Contaminated by Pesticides Body portion

Area (m2)

Proportion (%)

Whole body

1.85a

100

0.065

b

3.5

0.082

b

4.4

0.121

b

6.5

0.011

b

0.6

Front of neck and “V” of chest

0.015

b

0.8

Total, unclothed

294

Face Hands Forearms Back of neck

15.8

a

Surface area of a man 180 cm high weighing 70 kg, according to Sendroy and Cecchini (1954), or 1.88 m2, according to Geigy Scientific Tables (Lentner, 1984). b Values from Durham and Wolfe (1962); the calculated proportions are similar to those of Berkow (1931).

59.5.2.5  Use of Dyes as Indicators of Exposure As early as 1951, investigators at the Wenatchee Field Station of the U.S. Public Health Service explored use of a dye as an indicator of the drift of sprays. It was reasoned that exposure could be visualized directly and that quantitation could be achieved by simple colorimetry without the necessity of more complicated analytical chemistry. However, it soon was found by chemical analysis of residues that the dye and the insecticide for which it was intended as a surrogate did not behave in identical ways, and this approach was abandoned even though it did have the advantage of making contamination obvious. Others (e.g., Fenske et al., 1985) have investigated the use of dyes including some not susceptible to simple colorimetry but requiring computer-assisted video imaging for their visualization. However, the conclusions were the same: The visualization of exposure allows workers to participate in their own evaluation, but pesticides and tracer compounds penetrate fabric at different rates and a substantial proportion of tracer was lost during the sampling period.

59.5.3  Measurement of Oral Exposure It is common to warn workers against smoking or eating during work, lest they ingest pesticides transferred from their hands or clothing. Another more certain source of oral exposure involves blowing out clogged spray nozzles or the like. Although it is only common sense to avoid oral exposure, few studies have been made to learn how much real danger is involved. The analysis of vomitus or of stomach washings gives a minimal measure of the amount of toxicant swallowed. Lavage may be done experimentally as well as therapeutically.

Chapter | 59  Studies in Humans

However, the ordinary worker cannot be expected to cooperate by permitting gastric lavage merely for study. Durham and Wolfe (1962) proposed that methylene blue be combined with a formulation and that urinary excretion of the dye be used as a measure of oral ingestion inasmuch as dermal absorption of the dye is minimal. Apparently, the method has not been tested. Another method suggested for measurement of oral exposure involves analysis of food or objects that commonly are placed in the mouth after these things have been handled by formulators or sprayers while at work. Thus, Wolfe et al. (1963) asked workers to remove cigarettes from the package and handle them in the usual way in preparation for smoking. In some instances the worker also smoked the cigarette halfway. The unsmoked or partially smoked cigarettes were then protected from further contamination and taken to the laboratory for analysis. It was assumed that the amount of pesticide in the cigarettes might have been ingested or inhaled, depending on whether the contamination was on a portion that would have touched the lips (unsmoked cigarettes) or was on a portion that ordinarily would have been smoked (all cigarettes). In a similar way, sprayers were provided with sandwiches and other box-lunch foods that they carried into the field. After a predetermined period, each worker unwrapped the food and ate half of it. The other half, with which the worker’s hand had come in contact, was returned to the laboratory for analysis (Armstrong et al., 1973).

59.5.4  Problems of Measuring Separate Contributions from Different Routes of Exposure When comparing the importance of different routes of exposure, one can, of course, use the results of independent measurements of exposure by each route during the same operation. However, direct measures of exposure fail to take absorptive ability into account. It is possible that heavy contamination that is little absorbed by one route is less dangerous than light contamination that is efficiently absorbed by another route. Durham and Wolfe (1963) called attention to two special methods that may be of use when a question arises regarding the relative amount of toxicant absorbed by two routes. Both methods depend on measuring absorption in terms of urinary excretion of the toxicant or one or more of its metabolites. As described, both methods produce minimal measures of absorption by the dermal route.

59.5.4.1  Smyth Technique The first of these special methods apparently was first published by Durham and Wolfe (1963), but they attributed it to Dr. Henry Smyth, Jr. and spoke of it as the Smyth technique. The method consists of measuring the excretion of a

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toxicant or one of its metabolites and comparing the result with respiratory exposure estimated on the basis of respiratory volume and the measured concentration of toxicant in the air. If the amount found in the urine after a day, or some other definite unit of exposure, is fully accounted for by the amount of toxicant that would have been inhaled during the corresponding period of work, it simply is concluded that respiratory exposure accounts for all absorption that can be measured within the accuracy of the method. The same conclusion is reached if the amount of toxicant or metabolite excreted is less than the amount that would have been inhaled during the test interval. However, if the amount excreted is greater than can be accounted for by respiratory exposure, it is concluded that another route of exposure, usually the dermal, accounts for the difference. The method is not well adapted for measuring small differences of exposure, but it is a valuable technique for measuring larger differences. The method involves an assumption that oral exposure is trivial, but this assumption usually is valid. The method is limited by the efficiency of measuring excretion, which is why it gives only a minimal measure of exposure. The method is biased in favor of respiratory exposure because it assumes that all of the toxicant in the air would be inhaled regardless of particle size, and that all the inhaled material would reach the depths of the lungs and be retained. Actually, many particles in pesticidal aerosols and dusts are too large to be inhaled. Particles that are inhaled may impinge on the upper respiratory tract and eventually be swallowed. Gases, vapors, and extremely fine aerosols may be only partially retained. The special merit of the Smyth technique is that it permits the collection of useful quantitative data from measurements that may already be available or at least are easy to make. It can be carried out without any interference with the worker’s activity except that involved in the periodic collection of samples of urine.

59.5.4.2  Differential Protection Technique The second method of measuring the separate contribution of different routes of exposure referred to by Durham and Wolfe (1963) has been used by Dutkiewicz (1961) in his studies of aniline, but it still is not widely known. As a minimal measure of dermal absorption and as used by Dutkiewicz, the method can be considered a refinement of the Smyth technique, in which respiratory exposure and absorption are kept at zero by use of a supplied air respirator. Under these conditions, all of the excreted toxicant – including any that remains undetected – must have been absorbed from the skin. An extension of the latter method involves the use of a rubber suit, gloves, boots, and a complete plastic cover for the head with all potential openings sealed with tape so no air is permitted to enter, except through

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two small vents below the nose. In this form the method offers an accurate measure of respiratory absorption, provided the chosen measure of excretion is a good index of absorption. Unfortunately, this method of measuring respiratory absorption can be used only in certain circumstances. A person will be injured seriously if he stays long in the sun or in a hot place while sealed in rubber and plastic. Although the two forms of complete protection may be used separately, they can be employed to greatest advantage in experiments designed to exploit both, namely, what may be called the “complete differential protection technique.” Groups of workers carrying out essentially identical operations may be divided into three groups who receive complete respiratory, complete dermal, and no special protection, respectively. The latter group serves as a control and is given only such protection as ordinary clothing or whatever protection devices are considered standard for the operation under study. When exposed in this way for a work cycle or other predetermined period, the average absorption experienced by the controls should equal approximately the average absorption of one of the experimental groups plus that of the other. If absorption by one route is relatively very large, it may not be possible to distinguish the absorption in that group from absorption in the control.

59.5.5  Measurement of Absorbed Dose It would seem that determination of the absorbed dose is one of the most important measurements that could be made in any situation in which people encounter potentially toxic chemicals. An attempt often is made to estimate the ingested dose or the retained dose in those instances of acute poisoning in which any information on dose is available. In a very few of these cases the total amount of the compound or one of its metabolites recovered from a series of urines will have been measured so that one has an absolute measure of the minimal absorbed dose. Such information is no less important in toxicology than it is in clinical medicine, where the importance of information on dose is taken for granted and where information often is available on the degree of absorption of a particular drug when administered by a particular route. Such information on drugs is likely to be about equally available in connection with single and repeated doses. For the relatively few nontherapeutic xenobiotics that have been studied in volunteers, we have information on absorbed dose that is comparable to that for drugs with one crucial exception. These measurements for drugs involve pharmacologically active amounts; these measurements for nontherapeutic xenobiotics in volunteers almost always are restricted to doses too small to produce any clinical effect. The studies in volunteers are planned that way with good reason.

Hayes’ Handbook of Pesticide Toxicology

The largest number of persons substantially exposed to any nontherapeutic xenobiotic are workers. They are likely to greatly outnumber volunteers, and, if they do not greatly outnumber persons who are poisoned, the compound will be restricted or banned completely. At least until experience has accumulated, there can be no certainty that the doses that workers absorb are not pharmacologically active. There are at least three reasons for measuring the daily dose that workers absorb. First, for any given compound, its measured absorption by workers could be compared with the results of short-term and long-term studies in animals. For newly introduced compounds, such a comparison would offer the best evaluation of the safety of the workers unless comparable studies in volunteers were available already. Second, the accumulation of measurements of absorption of a wide range of compounds by workers coupled with knowledge of the long-term medical results of this absorption would offer the possibility of greater precision than now exists in extrapolating the results of animal experiments to humans. One reason for this is that the conditions under which animals are exposed often have little similarity to human exposure. Animals usually are exposed in a way that encourages maximal absorption of the administered dose. The same is not true of human exposure. Furthermore, people are more likely to be exposed by different routes simultaneously than are animals in experiments. Measurement of absorbed dose is the best way to avoid the inevitable differences between the exposure of humans and that of experimental animals and thus minimize the variables involved in comparing results in humans and animals. Finally, measurements of the dose absorbed by workers can be used as a point of reference in evaluating the effectiveness of the sum of all industrial hygiene procedures – whether factorywide or personal – in use at the time. What the worker absorbs is the ultimate measure of what penetrated the industrial hygiene net. Although total absence of absorption rarely is a practical goal, it is always the aesthetic goal. There is much to be said for a measure that indicates how nearly the ultimate goal is approached. The American Conference of Governmental Industrial Hygienists (ACGIH) (1988) now issues biological exposure indices (BEIS) as well as threshold limit values. It could be argued that those biological exposure indices that involve measurement of an excreted compound or metabolite (in contrast to a physiological change) are the equivalent of measurements of absorbed dose. To some extent this is true. Obviously, if working conditions remain the same, there will be some relationship between the daily absorbed dose and the concentration of a compound or a metabolite in a sample taken before a shift or at the end of a shift or whatever. However, it probably will be impossible to relate one to the other without a study of serial samples taken

Chapter | 59  Studies in Humans

over 24-h periods. For many compounds, the rapidity of excretion and other factors contributing to marked diurnal variation will preclude calculation of absorbed dose from the results of spot samples taken under specified conditions. It may well be that such spot samples evaluated according to biological exposure indices will remain the best way to monitor workers even for those compounds and situations in which the daily absorbed dose has been measured. However, the measurement of the daily absorbed dose under the same conditions will provide greater understanding of those conditions and a far greater possibility of meaningfully comparing different working conditions or of evaluating the safety of workers in terms of the results of studies in experimental animals. Of course the methods useful for analyzing a spot sample (Baselt, 1988) are equally applicable for measuring the total absorbed dose.

59.5.5.1  Practical Difficulties in Measuring Absorption Absorption by workers must be measured in terms of excretion. For practical purposes this is limited to urinary excretion for most compounds. Measurement of respiratory excretion may be practical for a few compounds. The difficulties associated with such measurements in urine include (a) imperfection of the method for analyzing a given compound (for example, extraction of samples may be incomplete), (b) the fact that true excretion may be not only urinary but also fecal and even (in the case of volatile compounds) respiratory, and (c) lack, for most compounds, of any study in which urinary excretion has been measured following doses known to be absorbed. In addition to similar difficulties associated with measurements of respiratory excretion, one must face the fact that it is difficult to collect samples and measure total expired volume simultaneously and that there is no such thing as a 24-h breath sample. All of these difficulties lead to the probability that, even with urinary samples, any determination of excretion will be a minimal indication of true absorption. However, in our present state of knowledge, even minimal measures of absorption under occupational conditions would be a great advance. Some further discussion of the difficulties listed above is given in the following section in connection with the few available examples of measurements of absorption.

59.5.5.2  Examples of Measurements of Absorption (a)  Absorption of DDT Table 59.5 lists the results of two studies to measure the daily dose of DDT absorbed by men engaged in its manufacture and formulation in different factories. These values for DDT are considered to be more precise than those

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Table 59.5  Absorption of Compounds by Workers under Ordinary Conditions Compound Absorption Kind of work (mg/worker/day)

Reference

DDT

14–42a

Manufacturing Ortelee and (1958); formulating Laws et al. (1967)

Parathion

0.2–1.2

Spraying orchards

Wolfe et al. (1970)

Carbaryl

5.6

Formulating

Corner et al. (1975)

2,4-Db

2.0–10.9

Spraying right-of-way

Libich et al. (1984)

Picloramb

0.10–0.46

Spraying right-of-way

Libich et al. (1984)

Aromatic aminesc

12.3–41.7d

Manufacturing Goldblatt (1949)

Benzidinee

0.4–0.9f

Manufacturing Ehrlicher (1958)

a

Range for workers with high exposure. 2,4-D and picloram were applied as a mixture (Tordon 101) in a ratio of 4:1. c Aromatic amines reported as -naphthylamine. d Average values for the whole factory for different years. e Benzidine based on analysis of its metabolite, monoacetylbenzidine. f Average values for different kinds of work. b

available for other compounds, because DDT is the only compound known to this author for which long-term studies in humans serve as a standard for measuring absorption in workers. Technical DDT was administered at rates of 3.5 and 35 mg/person/day to volunteers until they reached equilibrium of storage or at least approached it. The DDT was dissolved in peanut oil and presented in the form of a stable emulsion. Complete absorption of the doses was indicated by a failure to detect the compound or its metabolites in human feces and by animal experiments indicating that small dosages of dissolved DDT are absorbed completely. With few exceptions, the degree of urinary excretion of DDA measured in workers fell within the total range observed in volunteers who had received either 3.5 or 35 mg of technical DDT per person per day. A curve for average measured excretion at these two dosage levels served as a standard for estimating dosage (mg/person/day) for workers whose excretion of DDA had been measured. The potentially great difference between an estimate of absorbed dose based on a complete study in volunteers and an estimate based only on the amount measured in daily urine samples is illustrated by DDT. Four white men who received technical DDT at a rate of 35 mg/man/day showed no consistent change in their excretion of DDA

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after the 37th week of dosing until dosing was stopped at the end of the 94th week. Based on the reported excretion of DDA during this period of slightly over a year (Hayes et al., 1971), one can calculate that the average excretion of DDT-equivalent by these four men ranged from 2.88 to 8.15 mg/man/day and averaged 6.50 mg/man/day. Thus, measured excretion ranged from 8 to 23% and averaged only 19% of the known intake. (b)  Absorption of Parathion The study of parathion involved men operating air-blast spray machines to apply the compound to apple and pear orchards. The metabolite actually measured was p-nitrophenol, but results were expressed in terms of the active ingredient. The highest excretion rate observed was 0.168 mg/h, which, if maintained, would be equivalent to 4.03 mg/ man/day. However, excretion rose to a peak soon after each period of work but then fell rapidly. There was never more than one period of work in 1 day. When work was repeated on subsequent days, there was no evidence of increased excretion. As shown in Table 59.5, excretion – and, therefore, absorption – varied from 0.2 to 1.2 mg/man/ day for individual workers, and it averaged 0.53 mg/man/ day. These measures of absorption of parathion can be evaluated as accounting for over 86% of the total absorption of parathion, because (a) the study of each sprayman was based on an uninterrupted series of urine samples, (b) the analyses were continued for 4 or more days until excretion became undetectable, and (c) Morgan et al. (1977) were able to account, within 8 h of its ingestion, for 86% of doses of parathion administered to volunteers. All of the actual observations on parathion spraymen, including the highest rate of brief excretion, are consistent with the observation that volunteers in different studies tolerated daily oral doses of 7.2 mg/person/day without clinical effect and daily doses of 3.5, 4.5, and 4.8 mg/ person/day without effect on cholinesterase. (c)  Absorption of Carbaryl The study of carbaryl involved men working at the bagging and mixing stations in three formulating plants. Samples were collected serially so that 24-h urinary excretion was accounted for. Analysis was based on 1-naphthol and was done by two methods that were in good agreement. One of the methods involved acid hydrolysis and, therefore, would have detected the sum of conjugated and nonconjugated metabolites. However, there was no examination of possible fecal excretion and no background study of urinary excretion following doses known to be absorbed. Thus the observed excretion must be viewed as a minimal indication of absorption. (d)  Absorption of Herbicides The study of an herbicidal formulation involved a mixture of 2,4-D and picloram. The values recorded in

Hayes’ Handbook of Pesticide Toxicology

Table 59.5 are truly minimal. The ratio in which metabolites of the compounds were excreted (20:1) cannot be accounted for by the ratio in which they were applied (4:1). Possible explanations include lesser absorption of picloram or lesser recovery of picloram metabolites from urine, but any conclusion would require further study. (e)  Absorption of Carcinogens There are two reports of absorption of carcinogens or of compounds used as an indicator of one or more carcinogens in the organic intermediate and dye industry. The first report (Goldblatt, 1949) involved two factories in Britain where bladder tumors were observed among men exposed to a mixture of compounds and also among those thought to be exposed exclusively to aniline (unstated purity), benzidine, and a-naphthylamine containing 3–5% of the 13-isomer. For the latter group exposed to naphthylamines only, the average excretion of aromatic amines (reported as a-isomer) per worker per 1500 ml urine was reported for the whole plant, for different locations, and for different kinds of work. Excretion among those involved in flaking decreased from 74.8 to 69.5 to 43.6 mg/man/day during 1942, 1943, and 1944, respectively, as a result of improved working conditions. The least excretion in 1942 (24.0 mg/man/day) was found among those engaged in distillation. The average excretion for the whole plant decreased from 41.7 to 20.0 to 12.3 mg/man/day during the 3-year period. Using 4% as the proportion of 13-isomer in the mixture of isomers, the highest average exposure was 3.0 mg/man/day and the average for the entire plant in 1944 was 0.5 mg/man/day. Later, Ehrlicher (1958) reported excretion of monoacetylbenzidine by workers exposed to technical benzidine and no doubt other compounds. Apparently, benzidine itself and other metabolites of benzidine such as diacetylbenzidine were not measured. For most work situations, the results ranged from 0.40 to 0.90 mg/l in 1939, changed very little in 1951, and then decreased slowly and irregularly to 0.02–0.06 mg/l in 1958. If these values are multiplied by 1.5 (in consideration of the usual volume of urine excreted per day) the resulting range in 1939 was 0.6–1.3 mg/man/ day and that for 1958 was 0.03–0.09 mg/man/day.

59.5.5.3  Discussion The results summarized in Table 59.5 offer some idea of the degree of daily absorption of a few compounds under practical occupational conditions. Whereas some of the measures of excretion must be regarded as truly minimal estimates of absorption, there is reason to believe that the values for parathion are at least 86% of the true value and that the values for DDT are even better. The fact that the values for parathion are more than one order of magnitude less than those for DDT cannot be explained by any lesser tendency for absorption of parathion. On the contrary, when a mixture was applied, a

Chapter | 59  Studies in Humans

smaller proportion of DDT than of parathion was absorbed (Wolfe et al., 1970). The marked observed difference must depend, therefore, on the fact that exposure to parathion was under very different conditions than exposure to DDT, and greater care may have been taken with parathion because of its known danger. The recorded values for the excretion of the constituents of pesticide formulations fall within two orders of magnitude of one another. All of them represent substantial doses far in excess of the quantities of the same compounds to which the general public is exposed and presumably even more in excess of what the general public absorbs. Special interest attaches to the two studies of excretion under occupational conditions that led to bladder cancers. The report by Goldblatt (1949) indicates that, before conditions were proved, absorption of aromatic amines was substantial, being at least as great as that of DDT associated with its manufacture. Even when one considers the carcinogens themselves, measured average absorption was about as great as that for parathion and was of the same order of magnitude for 13-naphthylamine and benzidine. Neither Goldblatt (1949) nor Ehrlicher (1958) offered any information regarding the details of dosage response, but their reports do not exclude the possibility that tumors were more likely following above-average absorption. Unfortunately, both papers were written so soon after the improvement in working conditions that they document that there was no possibility of learning whether the decrease in average absorption led to any decrease in the incidence of tumors. Even the fragmentary information assembled here offers some idea of the dose of pesticides – and no doubt other chemicals – likely to be absorbed by workers. Unless specific measurements are available, the range of values already measured ought to be considered in estimating the probable absorption of any given compound and in judging the implications for workers of toxicity studies of that compound in animals. A comparison of the results of Ortelee (1958) and those of Laws et al. (1967) indicates that absorption of DDT in one factory was approximately double that in another. The results for carcinogens indicate that it was possible to reduce absorption by a factor of three or more by hygienic measures in the same factory. Thus, the use of measured absorption as a criterion for judging the value of specific changes in ventilation, personal hygiene, and the like was practiced as early as 1949. It is a pity that its use has not increased more rapidly.

59.6  Regulation of pesticides and other chemicals by the EPA The U.S. Environmental Protection Agency (U.S. EPA) regulates pesticides and other chemicals under a number of statutes for the protection of human health and the environment. Several laws require the EPA to make ­ decisions

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based on assessment of potential health effects, which requires collection of scientific data. Pesticides are regulated under the three statutes: (1) Federal Insecticide, Fungicide and Rodenticide Act (FIFRA);a (2) Federal Food, Drug and Cosmetic Act (FFDCA);b and (3) Food Quality Protection Act (FQPA)c of 1996. Under FIFRA, the EPA decides whether and under what conditions a pesticide can be used on food crops, without posing undue risks of adverse effects on human health or the environment. Under FFDCA, tolerances (maximal legally permissible levels) are established for pesticide residues in foods. FIFRA and FFDCA were substantially amended under provisions of FQPA. Periodic reevaluation of the registration (approval) and tolerance for each pesticide was required, to ensure that supporting scientific data remained up to date. Another important provision provided for an additional uncertainty (safety) factor (UF) of 10 to protect children and infants. This default factor is to be utilized in addition to intra- and interspecies UFs, unless there are data showing other values for one or more of these factors are more scientifically valid. Four federal laws place the burden of proof on third parties (entities other than the EPA) to conduct studies of adverse health effects of chemicals and to submit the findings to the EPA for evaluation. These statues include FIFRA, FFDCA, FQPA, and the Toxic Substances Control Act (TSCA)d. TSCA was enacted to give the EPA the ability to track and screen industrial chemicals produced by or imported into the United States. Toxicology testing can be required of compounds that may be a hazard to human health or the environment. Manufacture or import of compounds deemed to pose an unreasonable risk can be banned. Commercial sponsors of pesticides or potentially hazardous chemicals identified under TSCA must generally conduct extensive toxicity ­ testing in accordance with EPA guidelines (see Section 59.6.2). The burden of proof is on the EPA to assess health risks and to set standards for air and drinking water contaminants. The Clean Air Acte requires the EPA to significantly reduce daily emissions of pollutants known or suspected of causing serious health problems. The effectiveness of mandated control measures is to be assessed and additional standards adopted if necessary to mitigate remaining significant risksf. Under the Safe Drinking Water Act, the EPA sets standards to achieve levels of chemicals that maximize health risk reductionb benefits at a cost justified by the benefits. Industry can be required to conduct some tests and collect data, but the research burden is largely borne by the EPA. The agency conducts many investigations itself (intramural) and funds extramural research projects through contracts, a

Available at www.epa.gov/region5/defs/html/fifra.htm Available at www.fda.gov/opacom/laws/fdcact/fdctoc.htm c Available at www.epa.gov/oppfead/fqpa/backgrnd.htm d Available at www.epa.gov/region5/defs/html/tsca.htm e Available at www.epa.gov/air/oaqcaa.html f Available at www.epa.gov/air/criteria/html b

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cooperative agreements, and grants. This work sometimes involves human, as well as animal, research subjects.

59.6.1  EPA’ s Risk Assessment Approach EPA follows the standard four-step approach described by the NRC (1983) in assessing risks of pesticides and other chemicals to human health. Step 1 is termed Hazard Identification (HID). Its objective is to identify as many of a chemical’s adverse effects as possible. HID typically involves monitoring a large battery of tests in different species given high doses, often for a lifetime or even for successive generations. Step 2 entails Dose–Response Assessment. Test animals are usually given a series of doses orally and/or by inhalation, and sensitive toxicity parameters are monitored that were identified in the HID. Dose–effect (doses at which different adverse effects occur) and dose–response (association between dose and magnitude of a given effect) relationships can thereby be established. Step 3, Exposure Assessment, involves determining concentrations of chemical to which persons are exposed orally, by inhalation, and/or dermally. Risk Characterization (Step 4) is the final step in assessing human health risks. Risk depends upon both a compound’s toxicity and the likelihood of people coming into contact with it. The EPA establishes levels of pesticide exposure at which there is reasonable certainty of lack of noncancer effects for a lifetime of daily exposure. These acceptable exposure levels are termed the Reference Dose (RfD) for ingestion and the Reference Concentration (RfC) for inhalation. Calculations of RfDs and RfCs are usually based upon findings in dose–response studies in animals. For the sake of conservatism, the most sensitive effect in the most sensitive species is often selected as the “critical effect.” The basis for such a calculation is the highest no observed effect dose/level (NOEL) from the most appropriate study of adequate quality (i.e., the “principal study”). A lowest observed effect level (LOEL) can also be used as a starting point. On occasion the EPA utilizes a “benchmark dose” approach rather than UFs to derive RfDs. The EPA also develops cancer risk values for potential or known carcinogens, usually based on rodent data. Knowledge of qualitative and quantitative differences between animals and humans is often limited, necessitating the use of a UF to account for the possibility that humans are more susceptible. This is termed an interspecies UF. Values for this and other UFs generally range from 1 to 10. In the absence of applicable toxicity and/or toxicokinetic (TK) data for humans and the chosen laboratory animal, a full factor of 10 is typically utilized as a default. Substantial experimental evidence is required for adoption of lower values (Dourson et al., 1996). Most of the human study results submitted to the EPA after enactment of FQPA pertain to NOELs and/or LOELs.

Hayes’ Handbook of Pesticide Toxicology

NOELs/LOELs are routinely divided by additional UFs to obtain RfDs/RfCs that are deemed adequately protective of human health (U.S. EPA, 2002). The most commonly utilized is the intraspecies UF, which accounts for potential differences in susceptibility within the human population. A 10-fold default factor consists of a value of 3.2–4.0 for TK (i.e., absorption, distribution, metabolism, and elimination) differences and a value of 2.5–3.2 for toxicodynamic (TD) (i.e., sequence of events at the target site that result in toxicity) differences (Renwick, 1998). A number of other UFs may be evoked under given circumstances (Faustman and Omenn, 2008). If available toxicity data include a LOEL but no NOEL, the LOEL is divided by a factor of up to 10. Should there not be data from a chronic study, another factor of 10 is commonly applied. The FQPA, as described previously, specified inclusion of a UF of 10 for lack of knowledge of potential hazards to infants and children. Total UFs of up to 10,000 have been advocated in some circumstances, but an EPA panel proposed a limit of 3000 (U.S. EPA, 2002). Pesticide registration databases are usually so complete that calculation of acceptable daily intakes (RfDs) utilizes only interspecies, intraspecies, and child UFs.

59.6.2  Utility of Human Data in Risk Assessment Toxicity and TK data from human exposure studies make it possible for risk assessors to avoid many of the uncertainties inherent in interspecies extrapolations. Reliable human data are preferred by many regulatory agencies including the EPA (1994), FDA (2002), Health Canada (Meek et al., 1994), and the World Health Organization (IPCS, 1994). As the primary goal is to provide reasonable assurance of no harm from pesticide residues in foods, the most pertinent data are those (i.e., human data) that provide the greatest certainty that risk assessment-derived values accurately reflect human risks (McConnell, 2001). Knowledge of a pesticide’s toxic potential can be gained from several types of human studies. Much has been learned about harmful effects through case reports of accidental or intentional (e.g., suicidal) exposures, often to quite high doses. Investigations of workers exposed to pesticides during manufacturing and application processes also can yield important information. Occupational exposure levels generally exceed those encountered environmentally or resulting from food residues. The utility of epidemiological data from occupational studies, however, is usually limited by inadequate characterization of exposure to the chemical of interest and other chemicals, as well as by inability to recognize or control for confounding factors (Dourson et al., 2001). Most clinical case reports have the same limitations. Nevertheless, such data alert scientists as to particular toxic effects to anticipate for a

Chapter | 59  Studies in Humans

chemical and to use as “critical effects” to evaluate in subsequent research. Well-designed, ethical intentional dosing studies with humans can provide data that are very useful in evaluation of health effects. There may not be an appropriate animal model for certain chemically induced conditions (e.g., allergic reactions, exacerbation of respiratory problems in asthmatics by air pollutants). Intentional administration of such compounds to humans may be the only way to obtain data needed to set regulatory standards to protect public health. Experiments with humans typically involve giving low doses of drugs, vaccines, cosmetics, food additives, pesticides, or other chemicals. A useful human health effects study of a pesticide should include a range of doses, including at least one that is without effect and one affecting (a) sensitive biomarker(s). The effect(s) (e.g., slight inhibition of erythrocyte cholinesterase activity by a carbamate) should be mild and reversible, but ideally be linked to a serious, major consequence (e.g., pronounced neurotoxicity). In phase I clinical trials, potential therapeutic agents are given to some subjects in doses high enough to elicit side effects. This is routinely done to identify adverse effects to anticipate in medical practice. Compounds known or thought to be toxic (e.g., cancer chemotherapeutic agents) are administered to patient volunteers rather than healthy subjects (U.S. FDA, 2002). As the upper bound of dosages of pesticides and industrial chemicals is more limited in human studies, animals must be used to elucidate the full spectrum of adverse effects of test compounds. Experiments must also be conducted with animals to characterize chronic toxicity, carcinogenicity, and reproductive and developmental toxicity potential. Results from parallel human and animal experiments can substantially reduce uncertainties in extrapolating from animal studies to humans. Although amounts of pesticides that can be given to human subjects in vivo are quite limited, isolated human cells and tissues can be exposed to levels high enough to have a plethora of effects. Toxicity time and concentration dependency can readily be determined and compared in human and animal cells. These in vitro preparations can also be employed to study pesticide metabolism, identify toxic moieties, and elucidate mechanisms of injury in humans and test species (Gregus, 2008). Comparison of findings of TK and metabolism studies in human subjects and animals is another important means of extrapolating from one species to another (see below).

59.6.3  Attributes and Limitations of Laboratory Animal Data Assessment of the toxicity of pesticides and other chemicals in laboratory animals is the cornerstone of human safety evaluation. Defined studies are prescribed by the EPA for registration (i.e., approval) of pesticides

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(U.S. EPA, 1999). Cooper et al. (2006) have more recently advocated a tiered approach to assessing safety at different life stages. Animals are used for short, intermediate, and chronic studies, with which scientists can characterize a chemical’s spectrum of adverse effects over a wide range of doses and dosage regimens. More focused dose–response experiments employing a limited number of the more sensitive indices of injury can then be conducted. Most rodent species are inexpensive and easily maintained. Groups of uniform, well-characterized animals can be given precise doses of chemicals under carefully controlled conditions. Under these circumstances adverse effects of a single chemical or chemical mixture can be defined with greater certainty. In contrast, human populations are much more genetically diverse (Weber, 1999). There are also a wide variety of endogenous and exogenous factors that influence chemicals’ TK and toxicity (Dybing and Soderlund, 1999; Lof and Johanson, 1998). Many of these factors are not recognized or controlled for in human studies. Findings in animal toxicology studies are generally applicable to humans, although responses may differ quantitatively and/or qualitatively. The most definitive investigation to date was an in-depth review of data from preclinical (animal) and clinical (patient) studies of 150 compounds in 15 therapeutic classes (Olson et al., 2000). The data were supplied by 12 pharmaceutical companies. Interspecies concordance was assumed to exist if a compound had severe effects on the same organ. There was an overall concordance for 61% of the compounds. Interestingly, rodents were predictive for just 43% of the agents. In another comparative study, 43% of the clinical toxicities of 64 drugs in Japan were not forecast by animal experiments (Igarashi, 1994). The poorest concordance in this and the Olson et al. survey was for cutaneous hypersensitivity and for endocrine and liver dysfunctions. Obviously, assessments of animals cannot reveal subjective effects such as myalgias, headache, dizziness, nausea, or mental disturbances. Evaluations of susceptibilities to industrial and agricultural chemicals have provided additional information on the reliability of animal toxicity findings. Dourson et al. (2001) compared human data-based reference doses (RfDs) for 22 chemicals in EPA’s IRIS database with RfDs calculated from animal data using standard uncertainty factors. Seven of the 22 were insecticides, for which cholinesterase inhibition was measured in the human subjects. The overall interspecies concordance rate was 40%, with humanbased values lower for 32% of the 22 chemicals. Interspecies similarities and differences in susceptibility to chemical carcinogenesis have received considerable attention. Faustman and Omenn (2008) noted that all known human carcinogens adequately tested in rodents have caused cancer in at least one test species. Vinyl chloride, a widely recognized human liver carcinogen, caused liver tumors in mice and rats of both sexes. A review of National Toxicology Program bioassay data for 400 chemicals, however,

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showed only 23% of the animal carcinogens produced tumors in both mice and rats (Fung et al., 1995). Most such agents are apparently sex-, strain-, and/or species-specific (Grisham, 1996). Of necessity, rodents remain the animals of choice for assessments of pesticides’ and other chemicals’ carcinogenic potential in humans.

59.6.4  Use of Toxicokinetic Data for Species-to-Species Extrapolations Toxicokinetic (TK)-based conversions provide one of the most reliable means of extrapolating toxicology data from animal studies to humans (NRC, 1987). The fundamental goal of TK studies is to delineate the uptake and disposition of chemicals in the body. Toxicity is a dynamic process, in which the magnitude and duration of adverse effect are dictated by the concentration of the bioactive form of the chemical/moiety at its tissue target site and how long it remains there. These factors are determined by the net effect of the chemical’s systemic absorption, plasma protein binding, distribution and tissue uptake, biotransformation, interaction with cellular components, and elimination (Lehman-McKeeman, 2008). Metabolism may result in inactivation/detoxification (e.g., esterase-catalyzed hydrolysis of pyrethroids) and enhanced clearance, or in activation of the original chemical (i.e., parent compound) to a toxic metabolite (e.g., cytochrome P450-mediated sulfoxidation of organophosphates) (Parkinson and Ogilvie, 2008). It is important to establish by experimentation whether metabolism differs qualitatively and/or quantitatively in humans and the test species. Optimally, animal blood, target tissue, and urine time course data for the parent compound/bioactive moiety and major end metabolite(s) would be obtained for a wide dosage range, including nontoxic and toxic doses. Human data would necessarily be limited to low or trace doses of a pesticide. Computation of pertinent metabolic and TK parameters allows direct species-to-species comparisons of the compound’s fate and probable toxic actions. Physiologically based toxicokinetic (PBTK) modeling is the most precise and scientifically credible means of conducting route-to-route, high-to-low dose, and species-to-species extrapolations. PBTK models are enjoying increasing use for interspecies comparisons of drug (De Buck et al., 2007; Voisin et al., 1990) and chemical (Andersen, 2003) kinetics and toxicity. These computer models incorporate the unique anatomical, physiological, and metabolic characteristics of each species, as well as the compound’s physiochemical properties. Results of animal TK and metabolic experiments are utilized to develop and then to verify the model’s ability to accurately predict the time course of parent compound and/or bioactive metabolite in the animal’s blood and target tissue(s). The model is then allometrically scaled up to humans, or measured human physiological and

Hayes’ Handbook of Pesticide Toxicology

metabolic parameters are inputed. The human model is run to simulate blood and tissue time courses. Results of lowdose human experiments are compared with the computer simulations to validate the model (i.e., verify the accuracy of its predictions). A validated model allows one to determine the human exposure conditions that will produce target tissue levels of chemical equivalent to those previously found to cause or not to cause injury in the test animal. This approach has been used successfully for a number of pesticides including ethylene dibromide (Hissink et al., 2000), chlorpyrifos (Timchalk et al., 2002), and diazinon (Poet et al., 2004).

59.6.5  Events Leading to NRC’s Review of the EPA’s Use of Data from Human Dosing Studies Recommendations of the NRC (1993) Committee on Pesticides in the Diets of Infants and Children were incorporated into the Food Quality Protection Act (FQPA), leading to the EPA’s adoption of an additional 10-fold uncertainty factor (UF) to protect immature individuals (see Section 59.6.1). Three of the principal findings of the committee were the following: (1) most federal regulatory policies and practices did not specifically consider children; (2) few data were available on dietary intake of pesticides; and (3) very little information existed on potential health effects of pesticides or other chemicals in children. There was concern that developing organ systems (e.g., nervous, endocrine, immune, reproductive) might be particularly sensitive to some chemicals. There was also concern that physiological and biochemical changes during maturation could lead to toxicokinetic (TK) differences from adults (Bruckner, 2000). The new UF of 10 was recommended, largely due to the lack of studies employing sensitive measures of developmental effects. Any of the 10-fold default UFs previously described in Section 59.6.1 on the EPA’s risk assessment approach can be modified, if reliable data supporting another UF become available. The interspecies UF of 10 could be reduced to 1, for example, if findings in a reliable study demonstrate that the TK and toxicodynamics (TD) of a pesticide are comparable in adult humans and test animals. Conversely, a factor larger than 10 could be selected should humans be found to be much more susceptible. Adoption of the new/additional children’s UF raised concern among agricultural chemical companies that tolerances would be set so low that many commercially valuable pesticides could no longer be used. Results of 17 of 19 human dosing studies, furnished since 1991, were submitted by agrichemical companies to the EPA after enactment of the FQPA. Some advocacy and environmental groups objected, stating that such companies were subjecting research participants to potential risks, in order to

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offset FQPA’s new UF of 10 (i.e., provide human data to reduce the interspecies UF, in order to offset the children’s UF). Individuals who favored banning or reducing pesticide use argued that pesticide manufacturers have inherent conflicts of interest, in regards to the ethics of such human research and impartial interpretation of the data (Gorovitz and Robertson, 2000). Industry representatives and many members of the scientific community (McConnell, 2001; Resnik and Portier, 2005), however, believed that welldesigned, ethical human dosing studies can more accurately define safe exposure levels for compounds important in economical food production and in protection of public health. Continuing pressure from public interest groups caused EPA administrators to suspend use of human data in assessing pesticide risks in 1998, in order to consider scientific and ethical questions. The Environmental Working Group (1998), for example, published a report critical of intentional pesticide dosing experiments. The report recommended that the EPA impose a moratorium on human testing until the agency conducted an in-depth study of the process and adopted new policies. The EPA convened a joint subcommittee composed of members from its Science Advisory Board and FIFRA Scientific Advisory Committee. The majority of subcommittee members concluded that studies to safeguard human health could be scientifically and ethically acceptable if the information sought could not be obtained from other sources and if they met very high standards. It was recommended that (1) the EPA extend the protection of the Common Rule (40 EFR Part 26) to all human research activities; (2) actions of intra- and extramural internal review boards be carefully monitored by the EPA; (3) the EPA establish an internal ethics committee for compliance oversight; (4) earlier human data need not be rejected; and (5) there be special concern for the interests of vulnerable groups. Several subcommittee members wrote a minority report, in which they argued that risks to human subjects were downplayed, current design of studies was not scientifically rigorous, and the majority report would lead to more human testing and higher exposure of the general population to pesticides. Upon receipt of the joint subcommittee’s report, the EPA concluded that ethical and scientific questions remained and suspended use of human data from studies of all types of chemicals in December, 2001. The EPA made arrangements with the National Research Council (NRC) to form a committee to address these issues. Three agrichemical companies filed a motion in federal court, stating that the EPA’s moratorium constituted unlawful de facto regulation. After oral arguments in March, a U.S. Court of Appeals invalidated the EPA’s suspension directive in June, 2003. Participants in an independent workshop subsequently published a number of ethical and public policy recommendations on pesticide research in humans (Oleskey et al., 2004).

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59.6.6  EPA’s Charge to the NAS The EPA asked NRC to review and make recommendations on the complex ethical and scientific issues raised by the use of third party (i.e., a group other than the EPA) intentional human exposure studies of potential toxicants. It was requested that particular attention be paid to NOEL studies. The focus was not to be limited to pesticides. Guidance in a number of areas was requested, including (1) whether and how requirements of the Common Rule or other ethical standards should be applied to the conduct of third party toxicity research (TPTR); (2) how the submitter(s) of TPTR data should be required to document compliance with provisions for protection of human subjects; (3) whether there are alternative methods for obtaining data of comparable scientific merit; (4) what scientific standards should be established for reliability and relevance of results; (5) circumstances under which the EPA would consider removing or reducing the customary 10-fold interspecies UF; and (6) the manner by which the EPA should impose, implement, and enforce standards for conducting TPTR with human subjects.

59.6.7  NRC Committee and its Major Recommendations NRC formed the Committee on the Use of Third Party Toxicity Research with Human Research Participants to gather and analyze information pertinent to the tasks assigned by the EPA (see above) and to make recommendations for consideration in establishing agency policy. The committee was composed of members with expertise in toxicology, pharmacology, epidemiology, statistics, risk assessment, genetics, pediatrics, clinical trials, law, and ethics. The first meeting was held in December, 2002. The committee met six times over 12 months in open and closed sessions in which it invited testimony from a number of individuals, held a public forum to receive comments, and reviewed reports of human dosing studies of pesticides submitted by several companies. After careful evaluation and in-depth discussion of relevant information, the committee issued a comprehensive report (NRC, 2004) containing a number of recommendations to the EPA, institutional review boards (IRBs), research sponsors, and scientists. The NRC Committee recommended that intentional human dosing studies be conducted and used for EPA regulatory activities only if all of the following conditions are met: 1. The investigation addresses an important regulatory question that cannot be answered by animal studies or other types of human studies. The human dosing study is designed, conducted, and reported in a manner that assures it will be adequate scientifically to answer the question.

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2. The societal benefits outweigh any anticipated risks to participants. 3. There is a sound scientific basis (i.e., reasonable certainty) that participants will not be harmed. 4. All recognized ethical standards and procedures for protecting the interests of study subjects are observed, including equitable recruitment and selection of participants, informed consent, and independent review of the ethical and scientific merits of the investigation by an IRB or its foreign equivalent. The NRC Committee also recommended that the EPA establish its own advisory board to review both EPAsponsored and third party-funded human studies prior to and following the study. A primary purpose of this advice was to ensure in advance that an investigation was likely to yield scientifically valid results that would be useful for regulatory activities. The committee decided that a range of doses should be administered, including a NOEL and a LOEL for sensitive, reversible effects linked with adverse effects identified in previous animal, clinical, or epidemiological studies. Improving the accuracy of EPA regulatory decision making (e.g., establishing RfDs and RfCs) was judged to be a benefit to society. Human studies can generate information to use in protecting people from excessive exposures to chemicals other than pesticides. Thus, the committee’s charge was directed to human research conducted/funded by third parties or by the EPA. Seventeen recommendations were included in the NRC (2004) committee’s report. Their intent was to strengthen oversight and provide guidance for planning, conducting, and utilizing the findings of human toxicokinetic and toxicity research.

Conclusion Dr. Wayland J. Hayes was recognized internationally for his clinical expertise in the treatment of patients who were overexposed to pesticides and persons who were occupationally exposed to pesticides during their manufacture, formulation, and application. However, his recommendations and guidelines for the selection and treatment of volunteers used in studies to establish safe levels for regulating pesticide exposures in humans combined his clinical skills with an appreciation of the basic principles of toxicology. This combination together with his many years of hands-on experience in conducting and evaluating such studies have made his previous articles and reports on this topic a valuable resource. The most recent of his writings on this topic was his chapter in the first edition of this handbook (Hayes, and laws, 1991), and we have incorporated this material into this edition of the Hayes’ Handbook of Pesticide Toxicology. There have been many changes in the nearly two decades since Dr. Hayes defined the ethics and methodology for using volunteers to study

pesticides in humans, but the information in this chapter remains applicable and relevant for conducting such studies today. However, the EPA has recently addressed some of the issues and principles that were covered in the Hayes material, and there are now formal laws and regulations for some of these issues. There have also been significant developments in areas of toxicology and risk assessment that were not included in Dr. Hayes’ materials, and since many of these developments impact both the conduct and interpretation of studies in humans, we have added a section to this chapter to describe the history, purpose, and impact of these developments. We have reviewed the ethical and scientific issues raised by the use of third-party intentional human exposure to potential toxicants and have concluded this new section with some predictions regarding the future of using volunteers in pesticide research.

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Stumpf, S. E. (1967). Momentum and morality in medicine. Ann. Intern. Med. 67, 10–14. Su, M., Kinoshita, F. K., Frawley, J. P., and DuBois, K. P. (1971). Comparative inhibition of aliesterase and cholinesterase in rats fed eighteen organophosphorus insecticides. Toxicol. Appl. Pharmacol. 20, 241–249. Sun, M. (1981). Inmates sue to keep research in prisons. Science 212, 650–651. Taylor, C. (1941). Studies in exercise physiology. Am. J. Physiol. 135, 27–42. Tessari, J. D., and Spencer, D. L. (1971). Air sampling for pesticides in the human environment. J. Assoc. Off. Anal. Chem. 54, 1376–1382. Timchalk, C., Nolan, R. J., Mendrala, A. L., Dittenber, D. A., Brzak, K. A., and Mattsson, J. L. (2002). A physiologically based pharmacokinetic and pharmacodynamic (PBPK/PD) model for the organophosphate insecticide chlorpyrifos in rats and humans. Toxicol. Sci. 66, 34–53. Upholt, W. M., Quinby, G. E., Batchelor, G. S., and Thompson, J. P. (1956). Visual effects accompanying TEPP-induced miosis. AMA Arch. Ophthalmol. [N.S.J.] 56, 128–134. Voisin, E. M., Ruthsatz, M., Collins, J. M., and Hoyle, P. C. (1990). Extrapolation of animal toxicity to humans: Interspecies comparisons in drug development. Regul. Toxicol. Pharmacol. 12, 107–116. Weber, W. W. (1999). Populations and genetic polymorphisms. Mol. Diagnos. 4, 299–307. Wolfe, H. R., Durham, W. F., and Armstrong, J. F. (1963). Health hazards of the pesticides endnn and dieldrin. Hazards in some agricultural uses in the Pacific Northwest. Arch. Environ. Health 6, 458–464. Wolfe, H. R., Durham, W. F., and Armstrong, J. F. (1970). Urinary excretion of insecticide metabolites. Arch. Environ. Health 21, 711–716. World Health Organization (WHO) (1978). “Principles and Methods for Evaluating the Toxicity of Chemicals,” Part 1. Environmental Health Criteria 6. World Health Organ., Geneva. World Health Organization (WHO) (1982). “Field Survey of Exposure to Pesticides: Standard Protocol,” WHO Doe. VBC/82.1. World Health Organ., Geneva. [Reprinted in “Occupational Hazards of Pesticide Use” (G. J. Turnbull, ed.), App. 2, pp. 169–179. Taylor & Francis, London,1985; also in Toxicol. Lett. 33, 223–235 (1986)]. World Health Organization and the Council for International Organizations of Medical Sciences (WHO/CIOMS) (1982). “Proposed International Guidelines for Biomedical Research Involving Human Subjects.” World Health Organ./Counc. Int. Organ. Med. Sci., Geneva. World Medical Association (1964). Human experimentation. Br. Med. J. 2, 177. World Medical Association (1976). Declaration of Helsinki. Recom­ mendations Guiding Medical Doctors in Biomedical Research Involving Human Subjects. (Revised by the 29th World Medical Assembly, Tokyo, Japan, 1975). Med. J. Aust. 1, 206–207.

Chapter 60

Diagnosis and Treatment of Poisoning Due to Pesticides Wayne R. Snodgrass University of Texas Medical Branch, Galveston, Texas

60.1  Introduction Since the previous edition of this chapter in 2001, the principles of diagnosis, management, and treatment of poisoning by pesticides remain unchanged; there are increasing attempts to carry out prospective clinical studies to answer clinically significant challenges to improve the outcome of patients poisoned by pesticides. Pesticides have been used for many years to decrease adverse effects of a variety of pests. Pesticides may be categorized by the major agricultural classes of insecticides, herbicides, and fungicides. Additional groupings are rodenticides, nematocides, molluscides, acaricides, larvacides, miticides, pediculicides, scabicides, attractants (pheromones), defoliants, desiccants, plant growth regulators, and repellants. Approximately 1000 chemical compounds, biological agents, and physical agents, sometimes marketed as various brand names and formulations, are utilized in many areas of our planet Earth. In the United States, estimates are that about two-thirds of all agricultural utilization involves herbicides and about one-eighth involves insecticides; approximately 20% of usage is fungicides, fumigants, and other pesticides (U.S. EPA, 2001). The reader is referred to the chapter entitled “Diagnosis and Treatment of Poisoning” in the 1991 edition of this textbook for much useful and still relevant additional information. On a planet-wide basis, one estimate of human toxicity due to pesticides is up to 3 million cases per year of acute severe poisoning with perhaps an equal number of unreported cases and over 200,000 deaths per year (Ferrer et al., 1995; World Health Organization, 1990). In the United States, one estimate is 80,000 cases per year of pesticiderelated illness (Coye et al., 1986). Greater risk rates of pesticide poisonings may be expected in countries with less ­regulatory controls where pesticides may be used extensively. For example, about 13,000 hospital admissions and 1000 deaths associated with pesticide poisoning occur annually in Sri Lanka, which has a population of less than 15 million (Jeyaratnam, 1990). Recent data show that acute Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

pesticide poisoning in less developed countries continues to be a significant health problem (Murali et al., 2009).

60.2  Types of pesticides Organophosphates and N-methyl carbamates are the pesticides that most commonly cause systemic illness. Acute severe organophosphate poisoning is one of the most life-threatening human poisonings, but it is also treatable (atropine plus pralidoxime), often with a good outcome if treatment is begun promptly and early in the time course of poisoning. The organophosphates that cause the most illness in persons who do not work in agriculture are the moderate toxicity compounds chlorpyrifos, dichlorvos, dimethoate, malathion, and propetamphos, with some of these in recent years being less available to the general public in the United States. The organophosphates that cause the most illness in agricultural workers are the high toxicity compounds mevinphos, methomyl, methamidophos, oxydemeton, and parathion, and also the moderate toxicity compounds dimethoate and phosalone, with less agricultural use of some of these compounds in recent years in the United States. Carbamate pesticides typically cause less severe and shorter duration toxicity compared with organophosphates due to a more reversible complex formation between the carbamate and the cholinesterase enzyme (Fleming et al., 1997). Pyrethrin and synthetic pyrethroid insecticides cause much less systemic toxicity in humans compared with many other pesticides. Infrequent occupational poisoning and the rare occurrence of seizures have been reported (He et al., 1989). Extremely rare deaths (a total of two reported) are attributable to anaphylaxis from allergy. Organochlorine insecticides have low acute toxicity. DDT (dichlorodiphenyltrichloroethane) and chlordane are no longer used in the United States due to long-term environmental persistence. DDT also has a long-term persistence 1295

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in body fat; most people have low parts per billion (ppb) concentrations of DDE (1,1-dichloro-2,2-bis(p-chlorophenyl) ethylene), a metabolite of DDT, and low ppb levels of certain metabolites of chlordane (e.g., heptachlor epoxide) in their body fat. Endrin is more rapidly metabolized but is more acutely toxic. Food contamination by endrin has resulted in human illness with symptoms similar to encephalitis (Rowley et al., 1987). Chlorinated hydrocarbon insecticides are stable lipophilic chemicals and usually are contained in various organic solvents or as petroleum distillates. Often the petroleum distillates or organic solvents used as vehicles for the chemicals are as toxic as the pesticides themselves, and in the event of a significant ingestion, the vehicle toxicity should be considered as well (e.g., hydrocarbon pneumonitis). Many chlorinated hydrocarbon insecticides are rapidly absorbed and produce central nervous system (CNS) toxicity. Because of the halogenated nature of these organic compounds, hepatotoxicity, renal toxicity, and myocardial toxicity also may occur. Examples of chlorinated hydrocarbon pesticides include chlordane, DDT, dieldrin, chlordecone, lindane, toxaphene, and paradichlor­obenzene. Clinical manifestations after ingestion include apprehension, agitation, vomiting, gastrointestinal upset, abdominal pain, and CNS depression. Convulsions may occur at higher doses and may be preceded by symptoms of ataxia, muscle spasms, and fasciculations. In cases of ingestion, activated charcoal administration may be indicated if given early after ingestion. Epinephrine may be contraindicated because it may induce ventricular fibrillation as a result of sensitization of adrenergic receptors of the myocardium by halogenated hydrocarbons. Convulsions may be treated with midazolam followed by lorazepam in doses of 0.1–0.3 mg/kg administered intravenously over about 5 min. Methods to enhance elimination have yet not been successful other than as a supportive measure for hepatic and/or renal failure. Cholestyramine, which has been shown to bind chlordecone in the intestinal tract, may offer a means to treat chronic chlordecone poisoning and, pending further study, may have application to other agents (Boylan et al., 1978). Herbicides typically pose a low risk of acute systemic toxicity. The phenoxy herbicides [e.g., 2,4-D (2,4-dichlorophenoxy acetic acid)] may cause mild to ­moderate acute skin and pulmonary irritation. Peripheral neuropathy may occur after large skin exposures over a few days. Some evidence exists for a possible cancer risk with long-term occupational exposure (e.g., in farmers) to certain herbicides (Hoar et al., 1986; Zahm et al., 1997). Fungicides are a diverse group of compounds when grouped by chemical structure. Many fungicides do not pose a large risk of acute systemic toxicity; however, some are mutagenic for fungi and thus concern exists for possible chronic toxicity due to lingering traces in certain foods (Edwards et al., 1991).

60.3  General management of acute poisoning Clinical toxicology (also including the physician subspecialty of medical toxicology) is an area that encompasses both acute and chronic exposure to drugs, chemicals (including pesticides), and naturally occurring toxins. Human toxicological exposures range from acute overdoses (unintentional and deliberate) to chronic exposures (environmental and occupational). Treatment of a poisoned patient that is based on pharmacological principles promotes the use of rational methods that are beneficial to the patient’s recovery. There continues to be a great need for additional prospective, randomized, controlled, blind (if possible) clinical studies of treatment modalities in clinical toxicology. These are difficult studies to carry out in many instances. In the absence of such information, recommended therapies and procedures must be evaluated with appropriate skepticism and in the context of the best benefit-to-risk ratio for the individual patient. General management of pesticide poisoning is described in Table 60.1.

Table 60.1  General Management of Pesticide Poisonings Evaluate adequacy of oxygenation Ensure an open airway and adequate ventilation Assess and support vital signs Decontaminate and limit absorption Consider gastric lavage if within 1 or a very few hours of ingestion Consider administration of activated charcoal (should not contain sorbitol) Remove contaminated clothing and shoes Wash contaminated skin, eyes, hair Enhance elimination Consider administration of multiple dose activated charcoal (should not contain sorbitol) Supportive therapy Intravenous fluids and volume replacement Electrolyte balance Pulmonary support Blood pressure and cardiovascular support Antidote(s) if available Anticonvulsant drug therapy Close observation for 24 to 48 h after therapy stopped and recovery complete

Chapter | 60  Diagnosis and Treatment of Poisoning Due to Pesticides

Estimation of the severity of poisoning is an important initial step. A thorough history and details of exposure and/ or poisoning should be obtained, and a physical examination should be performed. Clinical evaluation includes the recommended use of one of several clinical scoring systems for coma and hyperactivity. These scoring systems serve as useful monitoring parameters to follow to determine whether the patient’s clinical status is improving or deteriorating. They also are useful to semiquantitate clinically the response to therapy. Table 60.2 shows an example of such scoring systems for coma and hyperactivity. The Glasgow Coma Scale, a more complex scale that is not shown, also is used widely to describe coma, although it does not take into account oxygenation status of the patient. In the United States and most developed countries there are now regional poison centers with expertise available for consultative toxicology services. These poison centers are available by telephone (800-222-1222 is the national phone number) to physicians and toxicologists who seek information from their large toxicology databases (Olson et al., 1991; Rumack, 2008). Lethal dose in 50% of animals tested (LD-50) and median lethal dose (MLD) determined typically in laboratory rodents have less value for quantitatively evaluating

Table 60.2  Scoring Systems for Coma and Hyperactivity Classification of coma 0

Asleep but can be aroused and can answer questions

1

Comatose; does withdraw from painful stimuli; reflexes intact

2

Comatose; does not withdraw from painful stimuli; most reflexes intact; no respiratory (ventilatory) or circulatory depression

3

Comatose; most or all reflexes are absent but without depression of ventilation or circulation

4

Comatose; reflexes absent; ventilatory depression with cyanosis, circulatory failure, or shock

Classification of hyperactivity 1

Restlessness, irritability, insomnia, tremor, hyperreflexia, sweating, mydriasis, flushing

2

Confusion, hyperactivity, hypertension, tachypnea, tachycardia, extrasystoles, sweating, mydriasis, flushing, mild hyperpyrexia

3

Delirium, mania, self-injury, marked hypertension, tachycardia, arrhythmias, hyperpyrexia

4

Above plus convulsions, coma, circulatory collapse

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clinical toxicity in humans. These parameters are not directly, or at times not even proportionally, extrapolatable to humans. The clinical history obtained in an overdose and/or exposure in an attempt to estimate the dose is known to be inaccurate in one-half or more of all cases. Thus, LD50 and MLD often are not used; instead, careful clinical monitoring along with the clinical history is done. A clinically useful estimate of toxicity is margin of safety (MS), which may be defined as LD-1 divided by ED-99, that is, the lethal dose (mg/kg) in 1% of a given human population divided by the therapeutically effective dose (for therapeutic drugs; mg/kg) in 99% of the population. This is a more conservative variation of the therapeutic index (TI  LD-50 divided by ED-50). The LD-1 often may be approximated from published case reports of overdoses or exposures. The ED-99 in the case of drugs may be estimated from therapeutic clinical study data of efficacy (dose–response curve). A drug with a high MS ratio value generally requires a considerably higher dose relative to the therapeutic dose to cause toxicity in a patient. Overall, the adage “treat the patient, not the poison” represents the most basic and important principle in clinical toxicology.

60.3.1  Skin Decontamination A major route of exposure to pesticides is by skin exposure and absorption. For some pesticides (e.g., some organophosphates) skin absorption may be so extensive as to result in severe poisoning. Due to the lipophilicity of many organophosphates, skin washing with an alcoholic (ethanol) solution is recommended in addition to washing with a detergent solution. Repeated skin washing (two, three, or more separate washings) with detergent solution as well as at least one or more alcoholic solution washings is recommended. Do not harshly scrub the skin, because skin abrasion perhaps may increase skin absorption of the pesticide. It is important to begin skin decontamination as soon as possible after the skin spill has occurred (Fredriksson, 1961; Wester and Maibach, 1985; Wolff et al., 1992).

60.3.2  Activated Charcoal Activated charcoal currently is the single most useful agent, when administered early, for prevention of absorption of orally ingested chemicals (including pesticides) and drugs. In addition, for other routes of exposure (e.g., skin and inhalation) for the many pesticides that likely undergo enterohepatic cycling or enteroenteric cycling, oral activated charcoal in multiple doses may be an effective adjunctive treatment due to its presumed ability to adsorb at least some pesticides and trap them in the intraluminal space of the intestine followed by rectal excretion; in effect, this provides a “sink” to trap and accelerate the removal of previously absorbed and distributed pesticide.

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Again, there is great need for prospective, randomized, controlled, blind (if possible) clinical studies to validate the efficacy of activated charcoal even for skin and inhalation routes of exposure, especially if enterohepatic and/or enteroenteric cycling is possible and even possible direct reverse transfer of a chemical from the mesenteric arterial supply toward the intraluminal intestinal compartment. In many cases of acute oral overdose, activated charcoal may be administered (typically 1.0 g/kg per orogastric tube of plain activated charcoal not containing sorbitol) without prior ipecac-induced emesis or gastric lavage. When the ingested agent is not adsorbed to charcoal or the situation warrants a different approach based on patient presentation and clinical judgment, other means of gastrointestinal decontamination may be considered (e.g., lavage or wholebowel irrigation). Prospective, randomized controlled clinical studies in orally drug-overdosed patients have indicated reproducibly that charcoal alone, compared with gastric emptying procedures with or without activated charcoal, either results similar patient outcomes with fewer complications or is superior (Albertson et al., 1989; Kulig et al., 1985; Merigian et al., 1990). There is a glaring lack of published data that detail the adsorption or nonadsorption of various pesticides to activated charcoal. There are some substances known not to be significantly adsorbed to activated charcoal. Relevant to pesticides, such substances include inorganic borates, inorganic bromides, and mineral acides and alkalais. There are some substances that at times are cited as not being adsorbed to activated charcoal, but for which there is published evidence that clinically significant adsorption occurs (i.e., 50 g of activated charcoal could adsorb an adult human toxic dose). Relevant to pesticides, such substances include carbamates, N-methyl carbamates, cyanide, DDT, diazinon, malathion, and mercuric chloride. There is great need for additional published comprehensive and detailed in vivo laboratory animal data [area under the curve (AUC), repeat plasma drug and/or chemical concentration measurements, and degree of change/ nonchange of LD-50 or LD-90] to determine the in vivo efficacy/nonefficacy of single-dose and multiple-dose activated charcoal for a variety of important clinically encountered pesticide inhalation, dermal, or oral exposures and/or overdoses. Pesticides for which there is a great need for in vivo laboratory animal data for activated charcoal efficacy/ nonefficacy include, but are not limited to, paraquat, chlorpyrifos, diazinon, organic mercurials, organic arsenicals, strychnine, lindane, and diethyl-m-toluamide (DEET). Multiple-dose activated charcoal (MDAC) involves the repeated administration (more than two doses) by oral or orogastric tube routes of activated charcoal to enhance the elimination of chemicals already absorbed into the body (typically 0.05–0.2 g/kg per hour continuous intragastric infusion for 1–3 days; use only plain activated charcoal; do not use combination products that also contain sorbitol).

Hayes’ Handbook of Pesticide Toxicology

Studies in animals and human volunteers have shown that MDAC increases chemical elimination significantly for a number of substances, but not all (e.g., substances with short half-lives and with extensive and high affinity plasma protein binding may show little enhancement of rate of elimination from the body). Few prospective randomized controlled clinical studies of MDAC in poisoned patients have been published. The most important endpoint to be measured is a possible reduction in morbidity and/or mortality. The early use of MDAC is an attractive alternative to more complex methods of enhancing toxin elimination, such as hemodialysis or hemoperfusion; however, the latter two modalities may be indicated in a relatively small subset of selected patients. Generally, patients with mild to moderate intoxications may benefit the most from MDAC. The decision to use MDAC depends on the physician’s clinical judgment regarding the expected clinical outcome, the efficacy of MDAC for the specific substance/toxin exposure, the presence of contraindications (e.g., intestinal obstruction) to the use of MDAC, and the effectiveness of alternative methods of therapy.

60.3.3  Gastric Lavage The efficacy with which gastric lavage removes gastric contents decreases with time after ingestion. Gastric lavage should be considered only if a patient has ingested a lifethreatening amount of a toxic agent, usually within 1 h or a very few hours previously. Gastric lavage with a large bore tube (36–40 French size in adult patients) inserted orally is a rapid way to remove some of the contents of the stomach; however, some contents may be pushed beyond the pylorus into the small intestine, increasing unwanted absorption of the chemical. There is no strong clinical evidence to support the opinion that lavage later than 1 h or a very few hours after a toxic substance ingestion will benefit patients; this typically includes ingested substances with anticholinergic activity and/or substances that delay gastric emptying. Nevertheless, uncommonly in individual patients, large amounts of gastric contents containing some of the ingested substance have been removed by gastric lavage many hours after ingestion, usually tablet formulations, not liquids or powders.

60.3.4  Cathartics The use of cathartics is no longer recommended routinely in the management of orally ingested poisons. There is no strong evidence to indicate that the use of cathartics improves patient outcome. Data suggest that cathartics may not alter patient outcome when adequate continuous intragastric activated charcoal is infused. When cathartics are used alone, the absorption of some substances (measured as AUC) is increased. Cathartics must not be administered

Chapter | 60  Diagnosis and Treatment of Poisoning Due to Pesticides

repeatedly to a patient with the absence of bowel sounds (i.e., absence of intestinal peristalsis). The uncommon pseudo-obstruction of the intestine during the administration of MDAC may not be prevented with the cathartic sorbitol (Longdon et al., 1992).

60.3.5  Whole-bowel Irrigation Whole-bowel irrigation (WBI) is not a routine procedure, but may be used in selected patients with a reasonable expectation of effectiveness (when measured as reduction of bioavailability) and relatively less evidence of adverse effects. WBI involves the enteral administration of large volumes of an isosmotic electrolyte lavage solution that contains poly(ethylene glycol) (PEG-ELS; e.g., Colyte or GoLytely) by orogastric tube at a rapid rate until the rectal effluent becomes clear. The purpose of this procedure is to irrigate out the contents of the gastrointestinal tract (GI) to prevent or decrease the absorption of toxic substances. The PEG-ELS is isosmotic and results in minimal or no detectable electrolyte and fluid changes in most patients. Published results typically show an approximately 65–75% reduction in bioavailability of ingested drugs (solid/tablet formulations) in volunteer studies. Adverse effects include vomiting from overly rapid infusion rates. Contraindications include ileus, GI hemorrhage, GI obstruction, and GI perforation. Relative contraindications include compromised circulation and a compromised airway. The WBI procedure consists of orogastric tube administration of PEG-ELS at a rate of about 25–40 mg/kg/h. The duration of infusion is determined by the goal of therapy, which may include passage of a clear rectal effluent.

60.3.6  Eye Contamination Absorption of liquid pesticide formulations may be very rapid from the eye or conjunctiva or the mucosa of the lacrimal duct and nose to which pesticides may drain from the eye. Especially lipophilic pesticides may absorb rapidly. Thus, begin as soon as possible to wash the eye with generous amounts of normal saline or water (normal saline is less irritating to the eyes). Use a gentle stream of saline or water. Saline or water may be poured into the eyes from a bottle or pitcher container. Eye irrigation for 15 min or more utilizing a clock to assure adequate time spent irrigating is recommended.

60.4  Acute poisoning by pesticides 60.4.1  Epidemiology Estimates of acute human planet-wide pesticide poisoning derived from mathematical models and projections range

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from 500,000 cases in 1972 to 25 million cases in 1990 (Levine and Doull, 1992). This compares with the 3 million cases cited at the beginning of this chapter. In 1985, an estimate of 1 million cases of acute pesticide poisoning and 200,000 deaths was accepted at a WHO informal meeting (Blondell, 1997; Levine and Doull, 1992). The WHO estimates that approximately 3% of agricultural workers in developing Asian countries have an episode of pesticide poisoning each year (Mohamed et al., 2008). When arranged by individual country, greater numbers of cases and deaths from pesticides occur in less developed countries; for example, approximately 90% of acute unintentional pesticide deaths occur in less developed countries. Approximately 99% of all deaths from pesticide poisoning (intentional plus unintentional) occur in developing countries (DeSilva et al., 2006). Regarding self-poisoning, one estimate is that 200,000 deaths from organophosphate selfpoisoning occur per year in rural regions of developing countries (Eddleston, 2008). Self-poisoning by use of pesticides accounts for about one-third of all suicides planetwide (Bertolote et al., 2006). During 1997–2002 in one district government hospital in India, 8040 patients were admitted with pesticide poisoning (Srinivas-Rao et al., 2005); 96% were intentional self-poisoning; the case fatality rate was 22.6%; two-thirds of patients were less than 30 years old; 57% were male; and monocrotophos and endosulfan accounted for the majority of deaths in which the pesticide was known (Srinivas-Rao et al., 2005). In the United States, 75% of pesticide usage is in agriculture (Calvert et al., 2008); in 3271 agricultural-associated cases of poisoning that were reviewed, 87% had a low severity of illness, 12% had medium severity, and 0.6% had high severity of poisoning; there was one fatal case out of the 3271 exposed persons (Calvert et al., 2008). Most acute pesticide poisonings in children occur in those who are less than 5 years old (Garry 2004). The wide range of overall estimates among studies occurs in part because of greatly varying assumptions made during calculations. Such calculated estimates often are made out of public necessity to assist public health officials in the absence of formal detailed accurate epidemiologic studies. Such estimates may not meet the usual epidemiologic standards of validity. Clearly, there is a need for better information. Regardless, the global magnitude of acute poisonings by pesticides continues to be enormous in costs of human suffering and human deaths.

60.4.2  Symptoms and Signs Diagnosis of acute organophosphate poisoning is based on a history of exposure, clinical symptoms and signs, and, where available, a blood test of red cell cholinesterase and plasma pseudocholinesterase. Tables 60.3 and 60.4 list symptoms and signs of cholinesterase-inhibitor poisoning.

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Table 60.3  Symptoms and Signs of Mild, Moderate, and Severe Cholinesterase-Inhibitor Poisoning

Table 60.4  Symptoms of Cholinesterase-Inhibitor Poisoning and Possible Drug Treatment

Exposure

Symptoms

Muscarinic (some may respond to atropine)

Mild

Anorexia Headache Dizziness Weakness Anxiety Tremors of tongue and eyelids Miosis

Moderate

Nausea Vomiting Salivation Tearing Abdominal cramps Diaphoresis Bradycardia Muscular fasciculations

Severe

Diarrhea Pinpoint and nonreactive pupils Pulmonary/ventilatory difficulty Pulmonary edema Cyanosis

  Excessive pulmonary tract secretions   Sweating   Salivation   Lacrimination   Miosis   Bradycardia   Hypotension   Urinary incontinence   Gastrointestinal spasms Nicotinic (may respond to pralidoxime)   Muscular fasciculations followed by weakness Central nervous system (some may respond to lorazepam/ diazepam)   Restlessness   Anxiety   Insomnia   Tremors   Convulsions   Ventilatory/pulmonary depression   Circulatory collapse

Loss of sphincter control Heart block Convulsions Coma, possible death

Table 60.5 lists a classification of organophosphate poisoning based on plasma pseudo-cholinesterase activity. Table 60.6 lists a few carbamate insecticides classified as moderately or highly toxic. The acronym MUDDLES, which stands for miosis, urination, diarrhea, diaphoresis, lacrimination, excitation of the central nervous system, and salivation, may be helpful in remembering the cholinergic excess signs that may occur to cholinesterase inhibitors. Another acronym, SLUD, which stands for salivation, lacrimation, urination, and defecation, also may assist in remembering cholinergic excess signs. However, both acronyms do not emphasize the most critical clinical signs that may be life threatening: the pulmonary signs of bronchorrhea (wet lungs) and bronchospasm, associated with bradycardia (usually). Severe central nervous system signs of coma and seizures also may occur in severe cases. Life-threatening apnea also may occur due to nicotinic receptor depolarization with resultant chest wall muscle paralysis; a fully atropinized patient may require mechanical ventilator support due to this phenomenon. It is crucial that treatment

Table 60.5  Classification of Organophosphate Poisoning Based on Plasma Pseudo-cholinesterase Activity Classification

Enzyme activity (% of normal)

Mild

20–50

Moderate

10–20

Severe

10

Note: Most patients who require pralidoxime will have a 50% or greater decrease in red blood cell (true) cholinesterase activity.

endpoints with antidotes emphasize monitoring of and improvement in bronchorrhea, bronchospasm, and bradycardia, and not miosis (small eye pupil size), because the latter is not life threatening. Perhaps a different acronym, BBB(u), which stands for bronchorrhea, bronchospasm, and bradycardia (usually), should be offered to assist in the assessment of the severity of acute organophosphate or carbamate poisoning. One study of scoring systems to predict outcome in acute organophosphate poisoning showed

Chapter | 60  Diagnosis and Treatment of Poisoning Due to Pesticides

Table 60.6  Toxicity Classification of a Few Carbamate Insecticides Moderately toxic   Bufencarb   Carbaryl   Methiocarb   Primicarb   Promecarb   Propoxur Highly toxic   Aldicarb   Aminocarb   Carbofuran   Dimetilan   Methomyl

no difference between the IPCS (International Program of Chemical Safety) PSS (Poison Severity Score) compared with the Glasgow Coma Scale (Davies et al., 2008); however, the identity of the organophosphate must be taken into account, for example, one-half of patients who died from fenthion poisoning had only mild symptoms/signs at the time of clinical presentation to the hospital (Davies et al., 2008). Exposure to organophosphates may produce a broad spectrum of clinical adverse effects that are indicative of massive overstimulation of the cholinergic system in the body. These adverse effects may present clinically as headache, weakness, dizziness, blurred vision, psychosis, respiratory (pulmonary) difficulty, paralysis, convulsions, and coma. A small percentage of patients may fail to demonstrate miosis, a classic diagnostic hallmark. Also, a few patients may not have bradycardia. The onset of clinical manifestations of organophosphate poisoning usually occurs within 12 h of exposure (typically shorter if intentional suicide ingestion). Measurement of red cell cholinesterase usually is diagnostic; when there is a reduction to 50% or less of control values, this indicates significant poisoning and the need to consider administration of pralidoxime (2-PAM) (a cholinesterase-regenerating agent) in addition to atropine. Efforts should be made to ensure that the patient does not become reexposed through contaminated clothing or reexposure to the contaminated environment. Children may show a different frequency of symptoms and signs compared with adults. In children, initial symptoms and signs of organophosphate poisoning more often include CNS depression, coma, stupor, dyspnea, and flaccidity, whereas miosis, excessive salivation, cold sweaty skin, and gastrointestinal symptoms may be less frequent

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(Sofer et al., 1989). Zweiner and Ginsburg (1988) reported the following pattern of clinical signs in infants and children: miosis (73%), excess salivation (70%), muscle weakness (68%), lethargy (54%), tachycardia (49%), seizures (22%), and pulmonary failure (38%). Young children may more frequently accidentally access pesticides stored in the home or garage (Reigart, 1995). Very limited data have been published regarding human reproductive/teratology risks for pesticides (Nurminen, 1995; Restrepo et al., 1990; Sever et al., 1997). Some symptoms and signs that occur with exposure to organophosphates may be due to impurities. For example, irritant (skin and upper airway tract) and odor adverse effects associated with organophosphates are likely to be due to low molecular weight mercaptans and sulfides (O’Malley, 1997). These mercaptans and sulfides also may cause headache and nausea; at times these substances also may result in bronchospasm. Thus, it may be necessary in a few cases to distinguish whether the bronchospasm is caused by the organophosphate or an impurity. An organophosphate-induced delayed neuropathy (OPIDN) may occur after an acute severe poisoning episode. Onset is about 7–21 days after exposure. Duration may be as short as many days or as long as months to years with full recovery or in some cases permanent impairment. Initial flaccidity with muscle weakness in the arms and legs that results in a clumsy shuffling gait is followed by spasticity, hypertonicity, hyperreflexia, clonus, and abnormal reflexes that indicate damage to the pyramidal tracts with a permanent upper motor neuron paralysis. In some patients, recovery occurs only in the arms and hands with no recovery in the lower extremities (foot drop, spasticity, and hyperactive reflexes), which is consistent with damage to the spinal cord (Ecobichon, 1996; Keifer et al., 1997; Lotti et al., 2005). OPIDN occurs primarily after a very large acute dose exposure, although it may occur after accumulated repeated lower doses. The intermediate syndrome is distinct from OPIDN in the following ways: onset within 24–96 h after recovery from acute cholinergic crisis, muscle weakness that primarily affects muscles innervated by the cranial nerves and proximal muscles, tetanic fade instead of denervation potentials on electromyography, and more rapid clinical recovery over 4–18 days compared with 6–12 months, as is typical in OPIDN. Persistent (several months), often mild CNS neuropsychological dysfunction may occur after an episode of acute organophosphate poisoning. Examples include persistence of decreased ability to concentrate on tasks, memory impairment, lethargy, emotional lability, blurred vision, muscle weakness, nausea, headaches, night sweats, and decreased performance (up to 2 years) on a World Health Organization (WHO) neuropsychological test battery (Rosenstock et al., 1991).

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60.4.3  Treatment The acute excess cholinergic symptoms and signs of organophosphate (and N-methyl carbamate) insecticides may be life threatening in severe cases. The main therapeutic emphasis in severe cases is to treat pulmonary and cardiovascular dysfunction and to treat with specific pharmacologic drugs [i.e., atropine and pralidoxime (2-PAM)]. Decontamination of the patient’s skin is another emphasis in cases of skin exposure. Oxygen and intravenous fluid therapy usually is needed in patients with severe cases of organophosphate poisoning. Mechanical ventilation also may be needed. Modern intensive care unit monitoring and therapy available in developed countries are indicated initially in most hospitalized patients. Clinical examination and arterial blood gas monitoring guides the use of oxygen and mechanical ventilation. Intensive care treatment more typically is required after deliberate or accidental ingestion of organophosphates than for occupational exposures. Atropine (which is antimuscarinic; no antinicotinic action) is the primary drug in the initial treatment of severe organophosphate poisoning. It is given in small but frequent repeated doses (e.g., 1–5 mg intravenously over 5 min, repeated at 15-min to several hour intervals; minimal pediatric doses are 0.05–0.1 mg/kg) and titrated to achieve the therapeutic goal of decreasing life-threatening excess cholinergic signs such as bronchorrhea (pulmonary rales and rhonchi heard on auscultation of the chest), bronchospasm (wheezes may be heard on auscultation of the chest), and bradycardia. Severely poisoned adults in one study required a mean of 23.4 mg (range 1–75 mg) of atropine intravenously to clear the lungs, to raise the pulse above 80 beats per minute, and to restore blood pressure to >�� 80 mmHg ������������������������������������������������ systolic; it was recommended that the most rapid atropinization was achieved with a regimen of increasing bolus doses of atropine after failure to respond to a previous bolus dose (Eddleston et al., 2004). A decrease in pulmonary secretions is evidence of clinical response once adequate doses have been administered. Doses of atropine may be titrated to maintain clear breath sounds and a heart rate (in adults) of 80–100 beats per minute. The total dose of atropine per day may markedly exceed that which is given for other purposes: for example, total doses of 1–30 g of atropine administered over several days have been reported in individual organophosphate poisoning cases (compared with the usual adult therapeutic dose of 1 or 2 mg). The endpoint of therapy is lack of symptoms in the absence of atropine dosing. Atropine should not be given until adequate ventilation is established and oxygenation has reversed hypoxia. Pralidoxime (2-PAM) is given in moderate and severe organophosphate poisoning cases. Its mechanism of action is to split apart the cholinesterase–organophosphate complex and regenerate active cholinesterase enzyme. It can

Hayes’ Handbook of Pesticide Toxicology

be effective against nicotinic, muscarinic, and CNS signs and symptoms. Pralidoxime has a short plasma half-life. It is perhaps best to administer it as a continuous infusion. One initial regimen suggested is a loading dose of pralidoxime of 4 mg/kg intravenously over 15 min followed by 3 mg/kg per hour continuous intravenous infusion. The dose is adjusted subsequently based on clinical response. Current recommendations are to not administer pralidoxime for a carbamate poisoning (Medicis et al., 1996; Tusk et al., 1997). Small preliminary studies suggest that magnesium sulfate may be of possible benefit in organophosphate poisoning treatment, but much more study is needed prior to recommending its routine administration (Eddleston, 2008). One interesting set of preliminary data showed that administration of fresh frozen plasma (FFP) (which contains cholinesterase enzyme activity) resulted in decreased mortality in organophosphate-poisoned patients. Of 33 patients with organophosphate poisoning, 12 patients were administered FFP in addition to atropine plus pralidoxime. Those patients receiving only atropine plus pralidoxime had a 14.3% mortality; those patients receiving FFP in addition to atropine plus pralidoxime had zero percent mortality (Guven et al., 2004). Further study is needed. Not all health care facilities in less developed countries may have FFP available; FFP administration carries the risk of transmission of hepatitis C and other viral diseases. Skin decontamination may be achieved by using soap washings followed by alcohol-soap washing with tincture of green soap or a similar mixture. Rescuers and medical personnel should be protected from contamination by using rubber gloves and aprons. A separate room in the emergency department of the hospital should be used if available and washing solutions should be appropriately discarded. For poisoning by phenoxyacid herbicides (e.g., 2,4-D), plasma kinetic evidence documents considerable shortening of plasma half-life by treatment with alkaline diuresis (see Table 60.7 for a list of some herbicides; see Table 60.8 for examples of chlorophenoxy herbicides). Benefit for clinical outcome perhaps might be expected. The risk of intravenous sodium bicarbonate fluid therapy and diuresis is relatively small compared with the toxicity of massive doses of phenoxyacid herbicides. Thus, alkaline diuresis therapy in selected cases of phenoxyacid herbicides appears to be therapeutically rational (Flanagan et al., 1990).

60.5  Chronic poisoning by pesticides The incidence and severity of chronic pesticide poisoning are unknown. Estimates of the prevalence of chronic pesticide poisoning apparently are not well known and not widely published (Maroni and Fait, 1993). One reason

Chapter | 60  Diagnosis and Treatment of Poisoning Due to Pesticides

Table 60.7  Herbicides: Some Common Classes and Names

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cigarette smoking. The production of organic pesticides is enormous: annual manufacture in the United States exceeds 1 billion pounds per year. There are approximately 1500 different ingredients in pesticides; 50,000 chemicals are in use (not including pesticides, pharmaceuticals, and food additives); there are about 4000 active ingredients in drugs; 2000 other compounds are used as excipients to promote stability and inhibit bacterial growth; 2500 additives are used for nutritional value and flavoring; and 3000 chemicals are used to prolong product shelf life (Maugh, 1978).

Class

Name

Acetanilides

Alachlor, Metolachlor

Amides

3,4-Dichloropropionanilide

Arylaliphatic acids

2-Methoxy-3,6-dichlorobenzoic acid

Carbamates

Isopropyl carbanilate

Dinitroanilines

a,a,a-Trifluor-2,6-dinitro-A, N-dipropylp-toluidine

Nitriles

2,6-Dichlorobenzonitrile

60.5.1  Types of Chronic Pesticide Toxicity

Substituted ureas

3-(p-Chlorophenyl)-1,1 -dimethylurea

Triazines

2-Chloro-4-(ethylamino)-6-(isopropy lamino)-S-triazine

The types of chronic pesticide toxicity may be categorized by chemical class (e.g., organophosphate or organochlo­ rine), by the presence or absence of current ongoing ex­posure, or by the extent of long-term body burden (e.g., a chemical with a very long total body elimination halftime measured in months to years). Organophosphate poisoning usually is considered to be relatively acute, for example, following termination of exposure, clinical toxicity resolves. However, clinical data clearly document adverse effects of organophosphates that persist long beyond the last period of exposure. Some data suggest that biological markers of increased susceptibility to organophosphates may exist, for example, serum ALA-D (aminolevulinic acid dehydratase) activity correlates with exposure, and PON1 (paraoxonase) and GST-T1 (glutathione transferase) appear to be related to susceptibility to chronic organophosphate poisoning (Hernandez et al., 2005). Having one of the two “slow metabolism” genotypes (alleles) of paraoxonase gene was a predictor of chronic toxicity symptoms in agriculture workers exposed to organophosphates [OR (odds ratio) 2.9], and there was a 75% prevalence of chronic toxicity symptoms/signs in pesticide applicators with the slow metabolism genotype (Lee et al., 2003). There are two general types of persistent organophosphate toxicity: delayed peripheral neurotoxicity and neuropsychological deficits. So-called neurotoxic esterase (NTE) or OPIDN poisoning is distinct from inhibition of plasma pseudocholinesterase or inhibition of erythrocyte cholinesterase. In the classic presentation of NTE-type toxicity, ataxic signs occur 1–3 weeks after an acute exposure. Painful transient paresthesias develop in a stocking-glove distribution over the lower extremities, rapidly extending to motor weakness and ataxia, with later involvement of the upper extremities. Evidence in laboratory animals points toward greater symptoms from multiple subchronic doses compared with a large single dose. Subchronic dermal exposure appears to be the most potent route of dosing. Structure–activity relationships of organic phosphonates versus phosphenates versus phosphates, chirality, and the ability to have a chemical leaving group correlate with ability to inhibit the NTE enzyme. Thus,

Table 60.8  Examples of Chlorophenoxy Herbicides 2,4-D (2,4-dichlorophenoxyacetic acid) 2,4-Dichlorophenoxybutyric acid Dichlorprop Silvex (2,4,5-trichlorophenoxy propionic acid) MCPA Mecoprop

for this is the difficulty of defining chronic pesticide toxicity in humans. Is it the incidence of neuropsychological deficits caused by chronic pesticide toxicity? If so, how do we measure such adverse effects objectively and quantitatively? Is it the incidence of chemically induced cancer caused by chronic pesticide exposure and long-term body burdens of certain long-lasting pesticides? Quantification of the risk for chemically induced cancer in humans is difficult in part because of the long (often 5–20 years or more), but unknown, lag time period for the cancer to appear clinically and the widely variable genetic susceptibility in different individuals. Despite these difficulties, there are certain principles and guidelines to assist in the evaluation of patients who claim to be or are suspected to be chronically exposed to pesticides. A large number and quantity of chemicals, including pesticides, result in a likely exposure risk for all individuals, especially if small and very small dose exposures are included. However, the magnitude of this risk for overall morbidity and mortality may be small compared with the risks of excessive alcoholic beverage intake and the risk of

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not all organophosphates pose a risk for NTE inhibition and only a few marketed organophosphates have been documented to cause clinical delayed neurotoxicity. A lymphocyte NTE test, if available, may help to assess exposure of an individual patient (Steenland et al., 1995). The second type of delayed organophosphate toxicity consists of neuropsychological deficits. Abnormalities of memory, abstraction, mood, intellectual functioning, and flexibility of thinking can be demonstrated in subjects with a last prior organophosphate poisoning of 7 years or more in a few cases. These individuals may have neuropsychological test results similar to subjects with cerebral damage or dysfunction. These sequelae are sufficiently subtle that clinical neurological examination, clinical EEG, audiometric, ophthalmic, and blood chemistry testing cannot discriminate previously poisoned subjects from controls. The physician cannot rely solely on the standard neurological examination or on clinical intuition to evaluate a patient who has been poisoned chronically by organophosphates (at lower doses that do not produce overt clinically toxicity). The usual clinical neurological examination does not thoroughly assess higher level cognitive skills, but rather focuses on sensory and motor functioning. The major deficits in these patients are cognitive and appear on neuropsychological tests of abilities that receive limited emphasis in the usual neurological examination. Thus, the two methods, clinical neurological examination and neuropsychological testing, provide a more thorough evaluation of such patients with chronic cognitive deficits following organophosphate poisoning (Eyer, 1995). Nerve conduction studies in acute organophosphate poisoning have been reported usually as normal. The muscle response to single stimulation shows a repetitive response. The repetitive response disappears on repeated stimulation, and this finding is thought by some to be characteristic. Organochlorine chronic pesticide poisoning presents equally difficult challenges for evaluation. Halogenated hydrocarbons include those used as pesticides as well as many fumigants and solvents. The fumigants and solvents, usually with molecular weights less than approximately 250 Daltons, generally are considered to have shorter halflives, in part due to their greater vapor pressures. The halogenated hydrocarbon insecticides, usually with molecular weights of approximately 290–550 Daltons, have lower vapor pressures and include lindane (gamma-BHC or gamma-benzene hexachloride), chlordane, heptachlor, methoxychlor, aldrin, mirex, and toxaphene. Other halogen­ ated hydrocarbon environmental contaminants also pose a toxicity risk for humans if there is a sufficiently high dose exposure, and include TCDD (dioxin or 2,3,7,8-tetrachlorodebenzo-p-dioxin), PBBs (polybrominated biphenyls), and PCBs (polychlorinated biphenyls). Many of these less volatile halogenated hydrocarbons share the properties of high lipid solubility, extensive distribution and storage in

Hayes’ Handbook of Pesticide Toxicology

body fat, and low rates and extents of metabolism to more polar metabolites with a resulting long total body half-life.

60.5.2  Background Accumulation An especially important factor in chronic pesticide poisoning is the potential for chronic persistent accumulation in the body. This accumulation occurs in all individuals because of nonoccupational exposure through food (even though the absolute amount may be very low), inhalation, or skin absorption. Thus, a background concentration (e.g., in body fat) of certain halogenated hydrocarbons, including PCBs, DDT, DDE (metabolite of DDT), hexachlorobenzene, dieldrin, lindane and its isomers, and sometimes TCDD (dioxin), pentachlorophenol, and BHT (butylhydroxytoluene), is detectable in the general human population. Also in body fat, detected more consistently in lower concentrations in nonoccupationally exposed individuals are chemicals such as aldrin, heptachlor, oxychlordane, mirex, chlrodecone, DDT, PBBs, chlorinated benzenes, polychlorinated terphenyls, tetrachlorophenol, toxaphene, and phthalate esters (e.g., DEHP, used as a plasticizer in plastics); detected less consistently are a variety of more volatile chlorinated aliphatic hydrocarbons. Thus, a definite body burden of halogenated hydrocarbons exists, a priori, in every patient. Subtle neuropsychological testing of absolutely unexposed individuals (absolute controls), including infants and children, does not exist; all individuals have some exposure to halogenated hydrocarbons. The potential for accumulation can be assessed by use of a bioconcentration factor (BCF). The BCF is the ratio between the concentration of a chemical in tissue (e.g., fat) and the concentration of the same chemical in the diet (both measured in mg/kg). A BCF of 1 or less means that no accumulation has occurred. The BCF for DDT plus metabolites is 1279; the BCF for hexachlorobenzene is 674; the BCF for lindane is 18. Some of these numbers are rather sobering. Such data are valuable in assessing relative risk for body accumulation. Another conceptually useful method for assessment of chronic toxicity, particularly in occupational exposure, is the chronic exposure index (CEI), CEI  log ( base 10)[ y  D divided by age minus 18]  1 where y is the number of years of exposure to a pesticide and D is the most recent estimate of number of days of usage of pesticides per year. Index values from the median value to the highest value for a particular group of subjects are defined as high chronic exposure; those from the lowest to the median value are called low chronic exposure. In addition, knowledge of relative chronic toxicity data also is useful in assessing overall risk, as is information on specific mechanisms such as metabolite-related tissue binding of halogenated hydrocarbons that result in more selective tissue toxicity.

Chapter | 60  Diagnosis and Treatment of Poisoning Due to Pesticides

60.5.3  Symptoms and Signs Symptoms and signs of chronic organochlorine and/or halogenated hydrocarbon pesticide poisoning are illustrated by the cyclodiene pesticides, which include chlordane, heptachlor, aldrin, and dieldrin. Two types of chronic exposure syndromes occur. In one, the exposure is continuous and leads to a slow accumulation of insecticide along with progressive symptoms. In the second type of chronic syndrome, insecticide exposure remains below that needed to cause symptoms; however, the individual experiences adverse effects with further intake. In 10–20% of sprayer operators who applied dieldrin and became poisoned, the earliest clinical signs developed in 3 months of exposure, but most required 8 months or more of exposure. Mild illness consisted of headache (that often was unresponsive to drugs and was persistent), dizziness, general malaise, insomnia, nausea, increased sweating, nystagmus, diplopia, tinnitus, slight involuntary movements, and blurred vision. Severe illness included progression to myoclonic jerking involving one or more limbs, sometimes accompanied by brief loss of consciousness. Approximately half of the 10–20% of cases progressed to convulsions. Persistent neurological sequelae from cyclodiene chronic poisoning include EEG abnormalities that last long after objective clinical signs of toxicity have resolved. Normalization of the EEG after exposure has stopped may require up to 3 months for endrin, a year for dieldrin, and more than a year for telodrin (Ecobichon and Joy, 1982). During continued chronic exposure to a mixture of chlorinated hydrocarbons, organophosphates, and carbamates in asymptomatic occupationally exposed workers, chronic elevation of serum epinephrine and of serum glucose (77 mg/ dl controls, 127 mg/dl exposed) occurred. Thus, neurologic effects of chronic pesticide exposure apparently include sympathetic stimulation of the adrenal glands. Similar data for altered neurological and altered dopamine and pituitary function are reported for exposure to the hydrocarbon styrene. Chronic poisoning from halogenated hydrocarbons results in measurable neurophysiological and neuropsychological abnormalities. Chronic toxic encephalopathy, once established, improves only slightly or not at all with time. Older individuals are more severely affected and less likely to recover. In one study, psychometric retesting 4 years later (4 years exposure-free) showed significant deterioration in verbal memory with improvement in visual memory. Computed tomography of the brain may or may not demonstrate loss of brain substance. Similarly, persistent long-term neuropsychological effects of TCDD, a halogenated hydrocarbon contaminate of 2,4-D but not itself a pesticide, have been studied. No clinically evident neurotoxicity was noted in an unreported study cited by Young (1984) of prison volunteers in the mid-1990s who were exposed dermally to 2,3,7,8-TCDD.

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The experiment was done to determine the dose of TCDD required to induce chloracne. A suspension of 1% TCDD in chloroform/ethanol was applied to the backs of 10 subjects on alternate days for 1 month. The cumulative dose applied was 7500 g of TCDD. Eight of the 10 subjects developed chloracne, which lasted 4–7 months. Blood chemistries including hematologic, liver, and kidney function remained within normal limits. TCDD is highly persistent with a whole body half-life of approximately 5 years. Poiger dosed himself orally with 105 ng of 3H-TCDD (tritium-TCDD) in corn oil. The dose of 1.14 ng/kg contained only 13 microCuries of tritium. Approximately 88.5% of the dose was absorbed. Radioactivity was detected only in the feces and was excreted at a rate of 0.03% of the body burden per day; the whole body half-time was estimated to be 5.8 years. Pharmacokinetic approaches to such long half-time chemicals continue to be developed. The Bhopal incident of massive methyl isocyanate exposure is an example of chronic chemical exposure. Methyl isocyanate is not a pesticide but does have a chemical structure that is similar to some pesticides. Although often considered a relatively short-term exposure, increasing evidence exists that the methyl isocyanate exposure at Bhopal has resulted in persistent long-term sequelae in survivors. The primary sequelae is chronic pulmonary damage, a combined obstructive and restrictive type. Methyl isocyanate-specific antibodies have been detected in the serum of the exposed persons upon follow-up and the antibody titer seems to correlate with the severity of the lung injury. Of over 200,000 people exposed to methyl isocyanate at Bhopal, estimates are that up to 5000 died within 2 days and about 60,000 individuals require long-term medical management.

60.5.4  Assessment of Exposure History One of the most difficult aspects in evaluating a patient who presents with or claims to have chronic pesticide poisoning is obtaining a meaningful medical history. Individuals with legitimate medical toxicologic events may be unable to reconstruct a completely useful history despite skillful questioning. Those individuals who present initially and have already decided that they will pursue legal redress for their exposure often have attributed (sometimes unknowingly) many minor and nonspecific health complaints to the alleged exposure. In an attempt to obtain a more consistent and thorough history, utilization of somewhat standardized clinical toxicology patient history and physical examination forms may be useful. The primary purpose of such a form is to assist in the completeness of the evaluation and, secondarily, to expedite the process. Such a form is particularly useful in the setting of an environmental–occupational toxicology clinic. The Agency for Toxic Substances Disease Registry has a useful form that is available.

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Risk analysis is another process the medical toxicologist enters into as he or she obtains an exposure history. Scientists have formulated various approaches to attempt to add precision and provide a more solid basis for governmental regulatory decisions regarding exposures to chemicals. However, for the physician evaluating an individual patient, obtaining as complete as possible the extent of exposure history and general knowledge of human toxicity effects for the individual chemical(s) usually is the most common historical information available for making a diagnosis and patient management recommendations. Equally important to chemical exposure information is information regarding the patient’s general medical status, past medical history, and any specific susceptibilities, for example known history of allergies or family history of cancer. Knowledge of the bioconcentration factor of the individual chemical(s) may be of use for selected chemicals. For infants, it is useful to know that breast-feeding infants potentially may be subjected to much higher halogenated hydrocarbon pesticide intake per kilogram of body weight than are adults due to the much higher concentration of some of these chemicals in human milk. Indeed, milk is the major route of excretion of body burdens of highly fat-soluble chemicals such as dioxin. To quote Matsumura (1985), “If a 5-kg infant consumed 0.7 liter of human milk daily containing DDT at an average concentration of 0.08 ppm, the resulting dose would be 0.0112 mg/kg/day. This value may be compared with the average adult daily intake of 0.0005 mg/kg/day. Assuming a two- to threefold increased sensitivity to the toxic effects of DDT in infants compared with adults, this amounts to eight times more than the FAO–WHO recommended maximum acceptable dose!” Similar considerations for lindane isomers (total benzene hexachloride) reveal that up to 1000 times more than the daily adult per kilogram intake may be “dosed” to infants (Matsumura, 1985). Subcutaneous fat concentrations of chlorinated hydrocarbons from infants and children sampled in 1982 reflected a highly significant association with the quantity of mother’s milk consumed. Some individual fat concentrations were higher than the mean values for adults from other areas. Because hydrocarbons have been documented to alter the developing brain biochemically in young laboratory animals, these human data have to be viewed with concern. Increasing data are available on quantitation of pesticide exposure in various situations. Home gardening ex­posure to carbamate carbaryl is approximately 8.5 mg per 10 g pesticide applied to garden plants. Protective clothing reduced exposure by 20-fold. Whereas over 90% of suburban residents in the United States are reported to use pesticides around the home, such exposure has considerable significance (Kurtz and Bode, 1985). Similarly, further published quantitative exposure data for aerial pesticide spraying would be useful.

Hayes’ Handbook of Pesticide Toxicology

60.5.5  Assessment of Symptoms Assessment of symptoms is a subjective evaluation. This is particularly difficult in chronic pesticide poisoning for at least two reasons: (1) many symptoms are nonspecific (e.g., headache, intermittent dizziness) or vague (e.g., general malaise); and (2) in the United States the patient may be planning legal action or be seeking workmen’s compensation relating to the pesticide exposure and thus may perceive his or her symptoms differently. It is important for the physician to exercise skillful judgment and, if possible, rely on prior experience. In the end, a judgment of the relevance of the complaints to the chemical exposure has to be made. The lay public’s perception of the relative risks of a variety of possible hazards may not correlate with actual relative risks (Baird, 1986). Interestingly, pesticides rank higher for greater perceived risk than, for example, motor vehicle crashes or alcoholic beverage-related risks. Physicians and scientists likely might rank certain hazards quite differently compared with the general public. However, such information about how the lay public perceives risk is useful in assessing which item and to what degree that item is likely to concern many patients. Risks taken voluntarily are about 100 times more acceptable than risks taken involuntarily (Baird, 1986).

60.5.6  Assessment of Signs Assessment of signs generally is an objective evaluation. In chronic pesticide poisoning, the physical examination should be done with particular attention toward target organs for toxicity of a particular chemical, if known. The neurological examination and close scrutiny for signs compatible with cancer are important when there is a long history of exposure. Biological monitoring is becoming increasingly sophisticated and helpful in the initial assessment of chronic pesticide exposure and for follow-up (He, 1993). Biological monitoring includes a variety of approaches, for example, direct measurement of blood, urine, or fat levels of a specific chemical, measurement of a metabolite in urine or of a toxicologically relevant metabolite (e.g., thioether compounds) in urine, or measurement of a biochemical effect relevant to toxicity risk (e.g., DNA adducts in urine; urine porphyrins). Unlike environmental monitoring (e.g., air level measurements of a specific chemical), biological monitoring can assess exposure by all routes, not just by the inhalation route. Biological monitoring has the potential to assess the actual uptake of a chemical by an individual. Also, biological monitoring, depending on the specific parameter, may take into account, at least in part, individual variability for risk from a given level of exposure. Biomonitoring of carbamate pesticides has been reported utilizing hemoglobin adducts (Sabbioni et al., 1990).

Chapter | 60  Diagnosis and Treatment of Poisoning Due to Pesticides

A biological exposure index (BEI) theoretically can be developed for many specific chemicals. For trichloroethylene (not a pesticide), the BEI is 100 mg trichloroacetic acid per liter of urine with the specimen collected at the end of the work week; that is, this is the acceptable upper limit of exposure, with greater concentrations indicating excessive exposure that requires further investigation (Lowry, 1987). One useful approach to biological monitoring for organophosphate pesticides, especially for low-level ex­posure, has been the repeated sequential measurement of plasma pseudocholinesterase activity to monitor organophosphate exposure of agricultural fieldworkers and those exposed to pesticide spray drift. Small but definite intraindividual plasma pseudocholinesterase differences over time were found in individuals who were intermittently exposed to organophosphate spray drift compared with those not exposed. Measurement of urinary alkylphosphates may be a more sensitive method for biological monitoring of low-level organophosphate exposure (Sunaga et al., 1989). Sequential monitoring for specific pesticides also appears to be indicated for indoor work areas such as greenhouses. Biological monitoring for risk of developing OPIDN (neurotoxic esterase inhibition) appears possible with the development of a lymphocyte and/or platelet NTE assay. Validation by careful clinical studies is needed prior to widespread application of this test. The porphyrin excretion patterns in urine may become a useful biological monitor for exposure to certain pesticides and chemicals. Compounds that are considered to be porphyrinogenic include some organophosphates and organochlorine pesticides as well as PCBs, PBBs, TCDD, vinyl chloride, and chlorinated naphthalenes (Strik, 1987). Chemically induced porphyria occurs due to inhibition of uroporphyrinogen decarboxylase, an enzyme that is part of heme synthesis. Inhibition of this enzyme results in accumulation and increased excretion of uroporphyrin and heptacarboxylic porphyrin. Only small elevations in urinary porphyrins have been seen generally in individuals exposed to chemicals such as PCBs. In one study, a few individuals following exposure to PCBs had a urinary uroporphyrin excretion of 66–106 g/24 h (Osterloh et al., 1987). Other studies have suggested that a total porphyrin (not uroporphyrin) excretion in adults of up to 200 g/liter of urine is normal (Strik, 1987). The pattern of porphyrins excreted rather than the total amount appears to be more significant for biological monitoring purposes. An increase in uroporphyrin and heptacarboxylic porphyrin may be an early finding in chemically induced porphyria. An increase in uroporphyrin and a decrease in coproporphyrin were seen following PCB exposure. Another potentially useful and rational approach to biological monitoring for selected pesticides and chemicals is the measurement of urinary thioethers. A usual but not necessary requirement is that the chemical be metabolically activated to an electrophilic (positively charged)

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short-lived toxic metabolite by the cytochrome P450 microsomal mixed function oxidase drug metabolizing enzymes. In addition to electrophilic intermediates, such chemicals as aliphatic and aromatic halides and alpha–beta unsaturated ketones may undergo direct (nonglutathione transferase) conjugation with glutathione. The technique exploits the fact that conjugation with glutathione, followed by urinary elimination as mercapturic metabolites, is a significant metabolic pathway, especially for putative toxic metabolites. Urinary thioethers potentially may indicate the extent of internal contamination of an exposed individual. Gasoline station attendants who dispensed gasoline directly (full-serve) compared with those attendants who did not (self-serve) had urinary thioether concentrations (7 pm to 8 pm time of single urine void) of approximately 8 mol–SH/mmol creatinine compared with approximately 2 mol–SH/mmol creatinine, respectively (Stock and Priestley, 1986). Although urinary thioether output detects dose-related increases in chemical exposure in animal models, this apparently has not been demonstrated in published clinical studies. Inhibition of red blood cell glutathione-S-transferase potentially may be another useful marker of toxic chemical body burden. Some organochlorine and halogenated hydrocarbon pesticides (many of which are no longer marketed in the United States) are known to be metabolically activated and are also known, upon chronic high dose exposure, to be carcinogenic in laboratory animals. Assessment of chronic exposure to these or similar pesticides by urinary thioether measurement may become a useful biologic monitor. Certainly this approach deserves further careful, well-designed, prospective clinical studies. A potentially highly significant biological monitor is DNA adduct formation. DNA adduct formation represents a direct chemical covalent bond between one of the bases of DNA and a chemical. The combination may be measured as a DNA adduct. Such adduct formation may be one of the initial steps necessary for some instances of chemically induced carcinogenesis. One has to be very cautious in attempting to apply such information clinically unless solid data validate a correlation between a given monitoring technique and the risk of human cancer. DNA adduct measurement in the white blood cells of iron foundry workers correlated quantitatively with air concentrations of polycyclic aromatic hydrocarbons such as benzpyrene (not a pesticide). The mean adduct levels (femtomoles of adduct per microgram of DNA) ranged from 0.24 to 1.5 fmol/g DNA (or, expressed differently, 0.80–5.0 mean adducts per 10 exponent 7 nucleotides) and correlated with air concentrations of 0.05 to 0.2 g of benzpyrene per cubic meter of air in the workplace. Similarly, vinyl chloride (not a pesticide), known to cause angiosarcoma of the liver in humans, reacts to form DNA adducts. One has to be cautious in interpreting such data. For example, although styrene (not a pesticide; used in the manufacture of plastics) produces chromosome aberrations

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in cultured human blood lymphocytes in vitro, workers exposed to low levels of styrene show no measurable chromosome abnormalities. Hemoglobin adducts of certain chemicals represent another biological monitoring endpoint and are described for ethylene oxide, benzene, benzpyrene, and dimethylnitrosamine (Farmer et al., 1987; Perera et al., 1988). Thus, many techniques exist or are being developed for biological monitoring. Some will prove to be more predictive than others. Biological monitoring of chronic pesticide exposure likely will expand as mechanisms of toxicity are elucidated further and new techniques advance.

60.5.7  Workup of Neurotoxicity The workup of neurotoxicity in individual patients can be divided into the assessment of the peripheral nervous system (PNS) and the central nervous system (CNS). In general, somewhat more objective measurements (neurophysiologic testing) can be made in examination of the PNS compared to the CNS (neuropsychologic testing) with the exception of objective CNS-evoked potential measurements. For PNS function, it is important to remember that axonal neuropathies often are associated with deficits of both sensory and motor function. Motor function testing includes: (1) inspection for muscle atrophy, unusual movements, and an analysis of coordination; (2) testing muscle tone and resistance to passive stretch of an extremity; (3) the Babinski reflex; and (4) analysis of the strength of individual muscles. Sensory function testing includes evoking the sensations produced by warmth and cold, pinprick, joint movement, tuning fork vibration, and shapes of complex objects. Cranial nerve examination, especially optic nerve (cranial nerve II) and trigeminal nerve (cranial nerve V) function, is important in evaluation of toxic exposures. Similarly, evaluation of the autonomic nervous system for bladder, bowel, and sexual functions, pupil response, lacrimation, salivation, sweating, and supine–upright blood pressure is important (Spencer et al., 1985). Knowledge of the specific toxin is helpful in planning and analyzing PNS evaluation. For example, in acrylamide (not a pesticide) neuropathy, sensory symptoms and signs are prominent. By contrast, NTE neuropathy from organophosphates may show retention of sensory function in the face of significant distal wasting and weakness. Nerve conduction velocity may not be altered in organophosphate NTE neuropathy, but a large drop in muscle action potential amplitude (EMG testing) may be observed. Acrylamide toxicity can be detected earlier by monitoring vibration sensation in the fingers (e.g., using a portable Optacon device; Spencer et al., 1985). In hexacarbon solvents (e.g., n-hexane or methylbutyl ketone; not pesticides) toxic neuropathy, use of nerve motor conduction velocity measurements is valuable because these solvents may slow

Hayes’ Handbook of Pesticide Toxicology

conduction. Blue-yellow color vision loss from solvent exposure is well documented. The Lanthony D-15 desaturated panel test for color vision loss has been validated for use at the workplace as an initial screening tool. Perhaps some of the foregoing testing procedures should be studied for their possible validity to monitor workers exposed to various pesticides. For CNS function, evaluation of mental status includes assessments of the level of consciousness, orientation, concentration, memory, cognitive functions, behavior, mood, and affect. The most frequently reported behavioral adverse effect of chemicals is a disturbance in psychomotor functioning. Usually, this is characterized by a delay or slowness in response time, clumsy or awkward eye–hand coordination or dexterity, or a combination of these. Diminished attention also has been found (Feldman et al., 1980). For a thorough evaluation of suspected neuropsychological deficits, use of a standard battery (group) of psychological tests is the best available procedure. Examples of such tests include the Halstead-Reitan battery, the Luria-Nebraska battery, and the Pittsburgh Occupational Exposure Test (POET) battery (Ryan et al., 1987). Formal testing is time-consuming and usually is done by psychologists trained in the use of such tests. There is a need for standardized neuropsychological tests that have a sufficient degree of simplicity and speed of administration such that they can be administered more readily to blue collar industrial workers as well as individuals with nonoccupational environmental exposures to pesticides. Furthermore, such test results must be adjusted for age and educational level (Feldman et al., 1980).

60.5.8  Risks of Pesticide-Induced Cancer Risks for cancer from occupational exposure to some chemicals, including pesticides, have been estimated. However, for a given individual, risk is nearly impossible to estimate with currently available data. Individuals vary widely in their genetic susceptibility, their exposures to other chemicals (e.g., cigarette smoke and ethanol), diet content, and other less well-defined parameters. Ecogenetics (distinct from pharmacogenetics) is the study of genetically controlled variations in response to environmental substances other than drugs. One example of the toxicological consequences of genetic differences is the 40-fold range among normal subjects in the capacity to metabolize chemical carcinogens by a specific cytochrome P450-dependent isozyme, aryl hydrocarbon hydroxylase (AHH) (Vesell, 1987). High activities are associated with lung carcinomas. It is likely that increased AHH activity would yield higher amounts of toxic carcinogenic reactive metabolites with exposure to polycyclic aromatic hydrocarbons (and perhaps pesticides with similar chemical structural groups). Such reactive metabolites bind covalently to DNA and may initiate the

Chapter | 60  Diagnosis and Treatment of Poisoning Due to Pesticides

development of cancer. Similar considerations apply to bladder cancer risk from aromatic amine exposure in the rubber industry. Some herbicides contain aromatic amines in their chemical structure. Numerous examples related to pesticides of suspected chemical-induced cancer exist. An increased risk for leukemia has been reported in children whose parents use pesticides in the home or the garden, or whose fathers are exposed occupationally to chlorinated solvents, or dyes or pigments (Lowengart et al., 1987); these data need independent confirmation. Phenoxy acid herbicide agriculture use has been reported to pose an increased risk for nonHodgkin’s lymphoma in farmers and forestry workers who use sprays (Zahm et al., 1997). Workers engaged in the storage and bulk handling of agricultural grains and peanuts in Sweden were reported to have an elevated risk for primary liver cancer; aflatoxin exposure was suspected. Use of insecticides and fumigants such as aluminum phosphide, carbon disulfide, carbon tetrachloride, ethylene dibromide, ethylene dichloride, malathion, and methyl bromide in flour milling may be implicated in the increased risk of non-Hodgkin’s lymphoma in flour mill workers.

60.5.9  Treatment Treatment of chronic pesticide poisoning or exposure requires careful medical evaluation and a thorough assessment of the exposure situation. The recommendation that the patient be removed from further exposure may be followed more easily for nonoccupational exposures. For occupational exposures, removal of the worker to another area of the worksite where lower exposure occurs may, on occasion, be acceptable medically. Alternatively, temporary removal while cleanup measures are begun also, on occasion, may be acceptable medically. Monitoring exposure at the worksite by a medically suitable means should be a part of all recommendations and follow-up evaluations of workers with chronic pesticide toxicity. Treatment of symptoms, other than removal from further exposure, often may be symptomatic only. Protective clothing can be recommended and should be worn for certain chemical exposure situations. More published data are becoming available to identify improved protective clothing materials, for example polyvinyl alcohol polymer material to resist permeation of methylene chloride (Stampfer et al., 1984). Gastrointestinal dialysis may offer the possibility of enhancing the rate of removal of the body burden of certain chemicals/pesticides that have long residence times (measured in months to years) and are sequestered in kinetically deep storage sites (usually body fat). A well-documented example was the use of repeated oral cholsestyramine therapy to increase the removal of chlordecone (Kepone), a highly neurotoxic chemical. Other examples include the

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use of repeated oral low-dose mineral oil for a relatively short time limit to enhance excretion of chlordane and lindane (Snodgrass et al., 1983, 1986).

Conclusion In summary, diagnosis and management of both acute and chronic pesticide poisoning present a significant challenge to current clinical capabilities. With careful attention to the details of exposure history and clinical findings in each individual patient, coupled with knowledge of the specific pesticide involved, an individualized management plan can be formulated that should be beneficial for most patients.

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Eyer, P. (1995). Neuropsychological changes by organophosphorus compounds: a review. Human Exper. Toxicol. 14, 857–864. Farmer, P. B. et al. (1987). Estimation of exposure of man to substances reacting covalently with macromolecules. Arch. Toxicol. 60, 251. Feldman, R. G. et al. (1980). Neuropsychological effects of industrial toxins: a review. Amer. J. Indus. Med. 1, 211–227. Ferrer, A. et al. (1995). Recent episodes of poisoning by pesticides. Toxicol. Lett. 82–83, 55–63. Flanagan, R. J. et al. (1990). Alkaline diuresis for acute poisoning with chlorophenoxy herbicides. Lancet 335, 454–458. Fleming, L. E. et al. (1997). Emerging issues in pesticide health studies. Occup. Med. 12, 387–397. Fredriksson, T. (1961). Percutaneous absorption of parathion and paraoxon. IV. Decontamination of human skin from parathion. Arch. Environ. Health 3, 185–188. Garry, V. F. (2004). Pesticides and children. Toxicol. Applied Pharmacol. 198, 152–163. Guven, M. et al. (2004). The effects of fresh frozen plasma on cholinesterase levels and outcomes in patients with organophosphate poisoning. Clin. Toxicol. 42, 617–623. He, F. (1993). Biological monitoring of occupational pesticide exposure. Int. Arch. Occup. Environ. Health 65(Suppl. 1), S69–S76. He, F. et al. (1989). Clinical manifestations and diagnosis of acute pyrethroid poisoning. Arch. Toxicol. 63, 54–58. Hernandez, A. F. et al. (2005). Changes in erythrocyte enzymes in humans long-term exposed to pesticides: influence of several markers of individual susceptibility. Toxicol. Letters 159, 13–21. Hoar, S. K. et al. (1986). Agricultural herbicide use and risk of lymphoma and soft tissue sarcoma. J. Amer. Med. Assoc. 256, 1141–1147. Jeyaratnam, J. (1990). Acute pesticide poisoning: a major global health problem. World Health Stat. Q. 43, 139–144. Keifer, M. C. et al. (1997). Chronic neurologic effects of pesticide overexposure. Occup. Med. 12, 291–304. Kulig, K. W. et al. (1985). Management of acutely poisoned patients without gastric emptying. Ann. Emerg. Med. 14, 562–567. Kurtz, D. A., and Bode, W. M. (1985). Application exposure to the home gardener. In “Dermal Exposure Related to Pesticide Use” (R. C. Honeycutt et al., eds.). American Chemical Society, Washington, DC. Lee, B. W. et al. (2003). Association between human paraoxonase gene polymorphism and chronic symptoms in pesticide-exposed workers. J. Occup. Environ. Med. 45, 118–122. Levine, R. S., and Doull, J. (1992). Global estimates of acute pesticide morbidity and mortality. Rev. Environ. Contam. Toxicol. 129, 29–50. Longdon, P. et al. (1992). Intestinal pseudo-obstruction following the use of enteral charcoal and sorbitol. Drug Saf. 7, 74–77. Lotti, M. et al. (2005). Organophosphate-induced delayed polyneuropathy. Toxicol. Rev. 24, 37–49. Lowengart, R. A. et al. (1987). Childhood leukemia and parents occupational and home exposure. J. Nat. Cancer Inst. 79, 39–46. Lowry, L. K. (1987). The biological exposure index: its use in assessing chemical exposures in the workplace. Toxicology 47, 55–69. Maroni, M., and Fait, A. (1993). Health effects in man from long-term exposure to pesticides. Toxicology 78, 1–174. Matsumura, F. (1985). Hazards to man and domestic animals. In “Toxicology of Insecticides” (F. Matsumura, ed.) 2nd ed. Plenum, New York. Maugh, T. H. (1978). Chemicals: how many are there?. Science 199, 162. Medicis, J. J. et al. (1996). Pharmacokinetics following a loading dose plus a continuous infusion of pralidoxime compared with the traditional

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short infusion regimen in human volunteers. J. Toxicol. Clin. Toxicol. 34, 289–295. Merigian, K. S. et al. (1990). Prospective evaluation of gastric emptying in self-poisoned patients. Amer. J. Emerg. Med. 8, 479–483. Mohamed, A. S. et al. (2008). Interethnic variability of plasma paraoxonase (PON1) activity towards organophosphates and PON1 polymorphisms among Asian populations. Ind. Health 46, 309–317. Murali, R. et al. (2009). Acute pesticide poisoning: 15 years experience in a large Northwest Indian hospital. Clin. Toxicol. 47, 35–38. Nurminen, T. (1995). Maternal pesticide exposure and pregnancy outcome. J. Occup. Med. 37, 935–940. Olson, D. K. et al. (1991). Pesticide poisoning surveillance through regional poison control centers. Amer. J. Public Health 81, 750–753. O’Malley, M. (1997). Clinical evaluation of pesticide exposure and poisonings. Lancet 349, 1161–1166. Osterloh, J. et al. (1987). Pilot survey of urinary porphyrins from persons transiently exposed to a PCB transformer fire. Clin. Toxicol. 24, 533. Perera, F. P. et al. (1988). Detection of polycyclic aromatic hydrocarbon DNA adducts in white blood cells of foundry workers. Cancer Res. 48, 2288. Reigart, J. R. (1995). Pesticides and children. Pediatric Ann. 24, 663–668. Restrepo, M. et al. (1990). Birth defects among children born to a population occupationally exposed to pesticides in Colombia. Scand. J. Work Environ. Health 16, 239–246. Rosenstock, L. et al. (1991). Chronic central nervous system effects of acute organophosphate pesticide intoxication. Lancet 338, 223–227. Rowley, D. L. et al. (1987). Convulsions caused by endrin poisoning in Pakistan. Pediatrics 79, 928–934. Rumack, B. H. (2008). “Poisindex,” Micromedex, Denver, CO. Ryan, C. M. et al. (1987). Assessment of neuropsychological dysfunction in the workplace: normative data from the Pittsburgh Occupational Exposures Test Battery. J. Clin. Exper. Neuropsychol. 9, 665–679. Sabbioni, G. et al. (1990). Biomonitoring of arylamines: hemoglobin adducts of urea and carbamate pesticides. Carcinogenesis 11, 111–115. Sever, L. E. et al. (1997). Reproductive and developmental effects of occupational pesticide exposure: the epidemiologic evidence. Occup. Med. 12, 305–325. Snodgrass, W. R. et al. (1983). Enhanced elimination of an environmental chlorinated hydrocarbon in man: use of oral mineral oil and cholestyramine. Veterin. Human Toxicol. 25(Suppl. 1), 59. Snodgrass, W. R. et al. (1986). Mobilization of a halogenated hydrocarbon pesticide from body fat in man: lindane. Veterin. Human Toxicol. 28, 471. Sofer, S. et al. (1989). Carbamate and organophosphate poisoning in early childhood. Pediatric Emerg. Care 5, 222–225. Spencer, P. S. et al. (1985). Chemicals causing disease of neurons and their processes. In “Neurotoxicity of Industrial and Commercial Chemicals” (J. L. O’Donoghue, ed.) Vol. 1, pp. 1–14. CRC Press, Boca Raton, FL. Srinivas-Rao, C. et al. (2005). Pesticide poisoning in south India: opportunities for prevention and improved medical management. Trop. Med. Internat. Health 10, 581–588. Stampfer, J. E. et al. (1984). Permeation of eleven protective garment materials by four organic solvents. Amer. Indust. Hyg. Assoc. J. 45, 642–654. Steenland, K. et al. (1995). Chronic neurological sequelae to organophosphate pesticide poisoning. Amer. J. Public Health 84, 731–736. Stock, J. K., and Priestley, B. G. (1986). Urinary thioether output as an index of occupational chemical exposure in petroleum retailers. Brit. J. Indust. Med. 43, 718.

Chapter | 60  Diagnosis and Treatment of Poisoning Due to Pesticides

Strik, J. J. T. (1987). Porphyrins in urine as an indication of exposure to chlorinated hydrocarbons. Ann. N.Y. Acad. Sci. 514, 219. Sunaga, M. et al. (1989). Urinary alkylphosphate levels as an index of exposure to organophosphate insecticides in pest control operators. Nippon Eiseigaku Zajsshi 44, 763–770. Tusk, G. M. et al. (1997). Pralidoxime continuous infusion in the treatment of organophosphate poisoning. Ann. Pharmacother. 31, 441–444. U.S. Environmental Protection Agency (2001). “Pesticide Industry Sales and Usage. 2000–2001 Market Estimates,” Office of Pesticide Programs, U.S. EPA, Washington DC. Vesell, E. S. (1987). Pharmacogenetic perspectives on susceptibility to toxic industrial chemicals. Brit. J. Indus. Med. 44, 505–509.

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Wester, R. D., and Maibach, H. I. (1985). In vivo percutaneous absorption and decontamination of pesticides in humans. J. Toxicol. Environ. Health 16, 25–37. Wolff, M. S. et al. (1992). Dermal levels of methylparathion, organochlorine pesticides, and acetylcholinesterase among formulators. Bull. Environ. Contam. Toxicol. 48, 671–678. World Health Organization (SHO) (1990). “Public Health Impact of Pesticides Used in Agriculture,” WHO, Geneva. Young, A. L. (1984). Determination and measurement of human exposure to the dibenzo-p-dioxins. Bull. Environ. Contam. Toxicol. 33, 702. Zahm, S. H. et al. (1997). Pesticides and cancer. Occup. Med. 12, 269–289. Zweiner, R. J., and Ginsburg, C. M. (1988). Organophosphate and carbamate poisoning in infants and children. Pediatrics 81, 121–126.

Chapter 61

Surveillance of Pesticide-Related Illness and Injury in Humans* Geoffrey M. Calvert1, Louise N. Mehler2, Judith Alsop3, Allison L. De Vries1, and Nida Besbelli4 1

Centers for Disease Control and Prevention, Cincinnati, Ohio California Environmental Protection Agency, Sacramento, California 3 California Poison Control System, Sacramento, California 4 World Health Organization, European Centre for Environment and Health, Bonn, Germany 2

61.1  Introduction A simple concise definition for public health surveillance is “data for action” (Giesecke, 1999). The traditional definition used by the Centers for Disease Control and Prevention (CDC) is the ongoing, systematic collection, analysis, and interpretation of outcome-specific data for use in the planning, implementation, and evaluation of public health practice (Thacker and Berkelman, 1988). There are three broad types of public health surveillance: hazard surveillance, exposure surveillance, and disease surveillance (Figure 61.1). Each of these types of surveillance can be used to generate data that are vital for targeting public health resources to prevent illnesses and injuries. Acute pesticide illnesses and injuries are preventable. The first step to preventing these illnesses and injuries is to understand the who, when, where, and why of these conditions. Public heath surveillance activities are vital to capturing this information (Stanbury et al., 2008). The most important use of surveillance data is to guide prevention activities, including regulatory, enforcement, consultative, or educational interventions. Surveillance can produce many data products that are useful for directing preventive action, including (1) estimation of the magnitude of the problem, (2) identification of trends in disease occurrence, (3) identification of epidemics or clusters of disease, (4) identification of emerging problems or new populations at risk of disease, and (5) evaluation of the effectiveness of

prevention and intervention efforts. Through the dissemination of these data, public health surveillance focuses attention on important health problems. The toxicity of pesticides continues to raise public concern and is the focus of much media attention. The importance of pesticides to protect the food supply and to control disease vectors is well recognized. However, it is also recognized that there is no perfectly safe form of pest control. Because society allows pesticides to be disseminated into the environment, society also incurs the obligation to track the health effects of pesticides. As such, surveillance of pesticide-related illness and injury continues to be important. Pesticide poisoning surveillance has been endorsed by several organizations, including the American Medical Association (1997), the Council of State and Territorial Epidemiologists (2009), the Pew Environmental Health Commission (2001), the National Institute for Occupational Safety and Health (NIOSH, 2001), and the U.S. Government Accountability Office (2000) (Figure 61.2).

Agent is a hazard Agent is present in environment

Hazard surveillance

Route of exposure exists Host is exposed to agent Agent reaches target tissue

Exposure surveillance

Agent produces adverse effect *

The findings and conclusions in this report are those of the authors and do not necessarily represent the views of the National Institute for Occupational Safety and Health or each author’s agency.

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Adverse effect becomes clinically apparent

Outcome surveillance

Figure 61.1  Three types of public health surveillance.

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Figure 61.2  A powered air-purifying respirator protects a worker using a thermal fogger in a greenhouse (courtesy of California EPA).

Pesticide poisoning among humans generally occurs either because of lack of compliance with existing pesticide regulations or because existing pesticide regulations are insufficient. The first cause involves cases that are preventable by following the precautionary measures specified on product labels and in governmental pesticide regulations. The appropriate interventions for these cases include enhanced education and enforcement. The second cause arises despite compliance with label instructions and regulatory measures and therefore requires interventions aimed at changing pesticide use practices and/or modifying regulatory measures. This chapter describes state-based, national, and international surveillance systems for pesticide-related illness and injury. Surveillance systems are the network of individuals and activities that engage in the process of surveillance. There are no national, comprehensive surveillance systems for pesticide-related illness or injury. Therefore, none of the surveillance systems described in this chapter provides a complete understanding of the pesticide-related illness problem. However, each system has strengths and weaknesses, and each system provides data that are useful for directing active intervention. The focus of this chapter is on surveillance systems that operate in the United States (both state based and national); however, some information is provided on international surveillance efforts. The chapter also describes some of the tools of surveillance (regulations that facilitate surveillance, efforts made to standardize case definitions and variables, and guidelines for evaluating surveillance systems). In addition, the chapter provides a general discussion of the limitations and strengths of surveillance data, with specific reference to the surveillance of pesticide-related illness and injury. Finally, the chapter provides an exploration of the role played by epidemiologic studies in the surveillance of pesticide-related illness and injury.

61.2  Surveillance systems 61.2.1  American Association of Poison Control Centers’ National Poison data System (a) Description Poison control centers (PCCs) are available free of charge to lay public and health professional callers 24 h a day, 7 days a week. A single toll-free number (800222-1222) was developed so all callers could access a poison center anywhere in the United States. By calling this toll-free number, the caller is electronically routed to the PCC serving his or her area. In 2007, PCCs managed 11,573 total poison calls a day in the United States, of which 6800 were human exposures cases (this equates to approximately one actual or suspected human exposure case every 12.7 s). This results in approximately 8.1 annual exposures per 1000 population nationwide (Bronstein et al., 2008). PCCs function primarily to provide poison information, telephone management and consultation, collect pertinent data, and deliver professional and public information. All 61 PCCs in the American Association of Poison Control Centers (AAPCC) participate in the National Poison Data System (NPDS) by providing their poison center call data. This data collection system was previously called the Toxic Exposure Surveillance System and was established in 1983. In 2006, NPDS was developed to create a real-time national poison exposure database and surveillance system. Sixty of the nation’s 61 poison centers (all except Puerto Rico) upload case data automatically every 1–60 min (mean time interval was 14 min in 2007), resulting in an almost real-time national exposure data collection and surveillance system (Bronstein et al., 2008). AAPCC was chartered in 1958 as a nonprofit, nongovernmental association to manage poisonings.

Chapter | 61  Surveillance of Pesticide-Related Illness and Injury in Humans

Table 61.1  Total Call Categories to Poison Control Centers, 2006–2007 Call category

Total

%

Total calls

8,257,436

100.0

Total human exposure cases

4,885,580

59.2

Total animal exposure cases

260,097

3.1

1,987,103

40.7

Non pharmaceutical exposures

2,451,521

50.2

Pesticide exposures

228,628

4.7

a

Pharmaceutical exposures

a

Confirmed nonexposure

2477

0.03

Total information cases

3,091,482

37.4

Pesticide information calls

8912

0.2

a

Single substance exposures only.

PCCs represent an independent member association that works closely with health departments and a variety of governmental agencies, including the Health Resources and Services Administration/Maternal and Child Health Bureau, Food and Drug Administration, U.S. Environmental Protection Agency (EPA), and the CDC (Bronstein et al., 2007). PCCs receive telephone calls from the public as well as health care professionals. Callers often call to request assistance to manage an exposure to a poison (62.3% of all calls). Another 37.4% of calls concern general information requests that do not involve a victim (e.g., drug interactions, drug identification, toxicity of plants, drug or chemical disposal, or medical information) (Table 61.1). Among all exposures, the proportion among children younger than 3 years is 38.1%, children aged 3–5 years 13%, children aged 6–12 years 6.3%, teens (aged 13–19 years) 7.0%, and adults (older than 19 years) 34.8%. Home is the most common site of exposure (91.2%), followed by the workplace (2.1%) and school (1.5%) (Bronstein et al., 2007, 2008). The route of exposure is most commonly by ingestion (77.8%), followed by dermal (7.4%), inhalation (5.7%), and ocular (5.1%). Of human exposures, 83.3% were unintentional, 12.9% were intentional, 2.5% were adverse reactions to products, and 1.2% included other reasons such as malicious intent, product contamination, or drug withdrawal. With advice from the PCC, 72.8% of cases are managed at home and 23.6% of cases are managed in a health care facility (HCF). The remainder (3.6%) are lost to follow-up (Bronstein et al., 2007, 2008). The age of the patient is associated with the treatment site, which is a surrogate for the severity of the exposure. HCF treatment occurs in 10.6% of exposed children

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younger than 6 years, 12.6% of exposed children aged 6–12 years, 47.4% of exposed teens (age 13–19 years), and 37.4% of exposed adults (older than 19 years). Young children are less likely to have severe illness because they are usually exposed to small amounts of substances that do not result in toxicity compared to teens and adults, who are more frequently exposed to larger amounts or to more dangerous substances (Bronstein et al., 2007, 2008). Poison calls are managed by specially trained health care professionals called “specialists in poison information” (SPIs) who are usually clinical pharmacists with a Doctor of Pharmacy (PharmD) degree or registered nurses. When the SPI has worked 2000 h and has managed 2000 human exposure cases, he or she is eligible to sit for the national certification examination given by AAPCC. After passing the certification examination, the SPI becomes a Certified Specialist in Poison Information (CSPI). Some poison centers also utilize adjunct health care professionals (e.g., pharmacy technicians, paramedics, and licensed vocational nurses) called “poison information providers” who manage less serious home cases and work under the overview of CSPIs. PCC managing directors provide direction, administrative supervision, and clinical education of staff. Most directors are board-certified clinical toxicologists who are Diplomates of the American Board of Applied Toxicology. Medical direction is provided by board-certified physician medical toxicologists who are available on call 24 h a day, 7 days a week. They provide medical oversight for staff and provide direction for the management of dangerous, difficult or unusual cases. Fifty-eight of the poison control centers (95%) submitting data to NPDS are certified PCCs. AAPCC developed certification criteria to recognize PCCs as qualified to serve their population. To be certified, a PCC must fulfill the following criteria to help ensure the quality of the center and the data (AAPCC, 2005): 1. Provide poison information 24 h a day, 365 days a year to both health professionals and the public. 2. Be readily accessible by telephone from all areas of the region. 3. Maintain comprehensive poison information resources. 4. Maintain written operational guidelines that provide a consistent approach to evaluation and management of toxic exposures. 5. Have a board-certified medical toxicologist on call at all times for medical direction. 6. Have a managing director who provides direct toxicologic supervision of PCC staff, strategic planning, and administrative oversight. 7. Have an SPI on duty at all times. SPIs must complete a training program and be certified. 8. Have an ongoing quality assurance program. 9. Keep standardized records on all cases handled by the center in a form that is acceptable as a medical record.

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10. Submit all human exposure data meeting quality requirements and all required data elements to the NPDS. Standard data elements include the time and date of the call, age and sex of the victim, location of victim at the time of exposure (e.g., home and workplace), location of the call (county, zip code, etc.), exposure substance, route of exposure, symptom assessment, source of treatment (e.g., on site or HCF), and an evaluation of medical outcome after case follow-up. Case follow-up generally involves telephone contact with the patient or the relevant health care professional. 11. Tabulate its experience for regional program evaluation and hazard surveillance. 12. Monitor the emergence of poisoning hazards and take action to eliminate poison hazards. 13. Provide information on the management of poisoning to health professionals throughout its region. 14. Provide a variety of public education activities, targeting “at-risk” populations. Poison center pesticide cases are identified through records of telephone calls from the lay public and health care professionals seeking information on how to manage an exposure to a pesticide. Poison exposure cases followed to a known outcome are categorized into outcomes of no effect, minor, moderate, major, death, or unrelated, depending on symptoms and severity. Definitions used by the PCCs to categorize medical outcome are summarized as follows (Bronstein et al., 2007, 2008): No effect: Patient developed no symptoms as a result of the exposure. Minor: Minimal symptoms that resolved quickly with no residual disability (e.g., mild gastrointestinal symptoms, skin irritation, and drowsiness). Moderate: Symptoms are more pronounced, prolonged, or more of a systemic nature than minor symptoms, result in no residual disability and are not life-threatening. Usually some treatment is indicated. Examples include corneal abrasions, high fever, disorientation, hypotension that rapidly responds to treatment, and isolated brief seizures. Major: Symptoms are life-threatening or result in residual disability or disfigurement. Examples include patients who require intubation plus mechanical ventilation and patients who sustain repeated seizures, cardiovascular instability, or coma. Death: Patient died as a result of the exposure or as a direct complication. Unrelated: The poison exposure most likely was not responsible for the effects. PCC SPIs rely on their experience and judgment to determine whether cases have symptoms consistent with the toxicology, dose, and timing of the pesticide exposure.

Hayes’ Handbook of Pesticide Toxicology

SPIs utilize information in PoisIndex, among other resources, to help them determine if symptoms are related or not. PoisIndex is a computerized database updated quarterly with information on approximately 1 million substances, including prescription, over-the-counter, generic, and foreign pharmaceuticals; street drugs; dietary supplements; cleaning agents; personal care items; household items; plants; insect, snake and animal bites; chemicals; pesticides; arts and crafts products; and industrial products (Klasco, 2009). Whereas 90.9% of cases involve exposure to a single substance, the remainder involve exposure to more than one substance or to unknown substances, making assessment of symptoms more difficult. Other situations that make assessment of a poison exposure more difficult and complex for PCC staff include the following: 1. The caller does not have the container at the time of the call and believes that the company or manufacturer name is the same as the product name. 2. The caller tries to describe the container over the phone with the belief that the description of the product will properly identify the exact product (e.g., “It’s in an orange can about 10 inches tall that you use to kill roaches”). 3. The supposed “inert ingredients” (i.e., the ingredients other than the active ingredients) are often not listed on the pesticide product label or in the PoisIndex product information. Not knowing if the pesticide product active ingredient is causing the symptoms rather than an unnamed surfactant, solvent, or petroleum distillate makes treatment decisions more difficult. If simple category information on inert ingredients such as “surfactants” were included on the label, medical treatment of exposures would be much easier. 4. The ongoing problem of placing pesticides in unlabeled food or drink storage containers. If an exposure to the unknown substance occurs in a situation such as this and no label information or other identifying information is available, valuable emergency treatment can be delayed, leading to major health consequences or even death. Patients treated at home or any other non-health care site are classified as “managed on site.” Patients seen in an HCF may be categorized as treated and released or admitted for medical care. “Admitted for medical care” is used when the patient is admitted as an inpatient to receive medical care. Some patients are admitted for further psychiatric evaluation rather than for medical care. For all human exposure cases seen at an HCF, 50.3% are treated and released, 14.9% are admitted to critical care, 8.7% are admitted for noncritical care, 1.9% are admitted for psychiatric evaluation, and 24% are lost to follow-up (Bronstein et al., 2007, 2008). In 2006, NPDS modified the way in which data are collected and interpreted, so this report uses data only for the

Chapter | 61  Surveillance of Pesticide-Related Illness and Injury in Humans

years 2006 and 2007. Beginning in 2006, data were collected only on single substance exposures rather than on multiple substance exposures, resulting in a better correlation among exposure, symptoms, and outcomes. In 2006, interpretation of death outcomes for poisoning fatality reports was also changed. In the past, some AAPCC death reports included cases such as a patient found dead surrounded by empty containers. No autopsy was performed, the cause of death was not obtained, and yet the case was coded as a fatality related to the products found by the body. In reality, death may or may not have been from exposure to the substances. Beginning in 2006, in an effort to make more evidence-based decisions on fatality cases, each death abstract was peer-reviewed by a team of two independent toxicologists consisting of a medical toxicologist and a clinical toxicologist. Case evidence was reviewed and ranked to determine if the exposure was the undoubted cause of death. Obtaining the medical examiner’s reports for each death case was encouraged, and the coroner’s results were also reviewed. In addition, a uniform method for collecting information on blood, plasma, serum, and vitreous toxin concentrations along with units and time of collection was also added. Using all the evidence gathered, a decision was made regarding whether the exposure was reasonably responsible or not responsible for the death. Only cases graded with reasonable confidence that the death was related to the exposure were reported as poisoning fatality cases. This resulted in smaller numbers of fatality cases being reported to the NPDS, but the cases were more authentically related to the exposure (Bronstein et al., 2007, 2008). PCCs were able to determine the medical outcome for 44.7% of all reported exposures. With outcome data, several measures of severity were used in this review of PCC data, including the percentage of cases reporting pesticiderelated symptoms or clinical effects; the percentage of cases that had a reported medical outcome, ranging from minor to death; and the percentage of pesticide exposure cases that were seen in an HCF. However, in many cases, patients presented to an HCF even though they did not have or never developed clinical symptoms. Fear and anxiety about the perceived risk of exposure may have played a role in the desire to seek medical attention. (b) Data Source All 61 of the nation’s PCCs participate in NPDS. Sixty of the 61 PCCs upload poison case data to NPDS automatically at a mean of every 14 min, resulting in near real-time surveillance and data collection. (Note that Puerto Rico did not have automatic case upload in 2006 and 2007.) (c) Target Population The entire U.S. population including all 50 states, the District of Columbia, American Samoa, Micronesia, Guam, Puerto Rico, and the American Virgin Islands is served by

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the PCCs included in NPDS. Because reporting of poison exposures is not mandatory, it must be kept in mind that PCCs do not collect all poison cases within their catchment areas. Of the 4 million poisoning cases estimated to occur in the United States annually (involving all poisons and not just pesticides), approximately 57% (2.27 million) are captured by the NPDS (Institute of Medicine, 2004). (d) Period of Data Collection All NPDS data for the years 2006 and 2007 were utilized. (e) Periodicity of Reports Data are available by calendar year. The NPDS annual report, which includes information on all types of poison exposures, is published annually in the December issue of Clinical Toxicology. Additional NPDS data can be purchased from the AAPCC (see www.aapcc.org). (f) Results As described previously, PCCs are also used as a resource on a variety of pesticide-related information questions. Table 61.2 provides the types of pesticide information calls received by PCCs in 2006 and 2007. The current review of pesticide exposure cases is based on 228,628 records of pesticide-related exposures reported to PCCs participating in NPDS during 2006 and 2007 (see Table 61.1). Pesticides accounted for 4.7% of all human exposures reported to PCCs. In contrast, for single substance exposures, 40.7% of all exposures were due to pharmaceuticals and 50.2% were due to exposure to nonpharmaceuticals. The leading cause of pesticide exposures was from insecticides (45.0%). Among insecticides, pyrethrins were the most frequently involved (10%) (Table 61.3). Although attempted suicide cases comprised 8.4% of all calls to PCCs, pesticides are used in only 1.7% of intentional poison exposures. Whereas pesticides ranked as the eighth most frequent category of exposures in both adult and pediatric populations, pesticides are not listed in the top 20 categories most frequently associated with the largest number of fatalities. The 10 leading categories of attempted suicide, as captured by PCCs, are sedative hypnotics/antipsychotics, opioids, antidepressants, acetaminophen combination products, cardiovascular drugs, stimulant street drugs, alcohols, acetaminophen as a sole ingredient, anticonvulsants, and fumes/gases/vapors such as carbon monoxide (Bronstein et al., 2007, 2008). A primary factor in determining risk to pesticides is the age of the patient (Table 61.4). Children younger than 6 years accounted for 87.90% of the anticoagulant rodenticide exposures, 83.8% of boric acid insecticide exposures, 65.5% of phenol-containing disinfectant exposures, and 51.9% of all disinfectant exposures. Unlike the other types of pesticides, rodenticides and roach baits are used almost exclusively as poison baits placed on the floor where

Hayes’ Handbook of Pesticide Toxicology

1318

Table 61.2  Information Calls Made to Poison Control Centers Explicitly or Implicitly Related to Pesticides, 2006–2007 Type of poison center information calls No. explicitly related to pesticides

%

General questions about pesticide 1648 applications made by a professional pest control officer

18.5

General pesticide information (other)

73.0

6505

Immediate referral to pesticide hotline because of suspected poisoning

759

Total

8912

8.5 100.0

Other types of poison center information calls that may apply to pesticides

Table 61.3  Pesticide Exposures by Pesticide Functional and Chemical Class, 2006–2007 Pesticide type

Total

%

Hypochlorite

22,013

9.6

Phenol

8579

3.8

Pine oil

5029

2.2

Other

11,841

5.2

Total

47,462

20.8

Sulfuryl fluoride

424

0.2

Other

340

1.5

Total

764

0.3

Fungicides

2260

1.0

Chlorphenoxy compounds

4039

1.8

Disinfectants

Fumigants

Safe disposal of chemicals

3979

20.0

Safe handling of workplace chemicals

298

1.5

Safe use of household products

8234

41.4

Routine toxicity monitoring

148

0.7

Clarification of media reports of environmental contamination

107

0.5

Glyphosate

8319

3.6

Water purity/contamination

2240

11.3

Other

4850

2.1

Potential toxicity of chemicals in the environment

3386

17.0%

Total

17,208

7.5

General questions about contamination of air and/or soil

1518

7.6

Boric acid/borates

8394

3.7

19,910

100.0

Carbamates

4898

2.1

Chlorinated hydrocarbons

1042

0.5

Organophosphates

7680

3.4

Pyrethrins

10,235

4.5

Others and combinations

70,577

30.0

Total

102,826

45.0

Insect, DEET

15,076

6.6

Moth

7706

3.4

Other

4452

1.9

Total

27,234

11.9

Anticoagulant

25,140

11.0

Other

5734

2.5

Total

30,874

13.5

228,628

100

Total

young children have access to them. Greater use of childresistant packaging would substantially reduce the number of pediatric cases. This is particularly true for disinfectants and other products stored in bottles. For baits, tamperresistant bait stations can be used to reduce exposures to children. Table 61.5 provides a comparison of substance exposures by age category. For DEET-containing insect repellents, 65.4% of cases involved children younger than 6 years. Many of these cases most likely occurred when the product was applied on them, resulting in the child sampling the bottle contents, licking exposed skin areas where the insecticide was applied, or rubbing exposed hands into his or her eyes. By contrast, only 14.1% of the fumigant cases and 21.7% of the fungicides involved children younger than 6 years. These pesticides are more likely to be used and stored outside or where young children have little access. Insecticide exposures were involved in 38.2% of cases in children younger than 6 years.

Herbicides

Insecticides

Repellants

Rodenticides

Total

Chapter | 61  Surveillance of Pesticide-Related Illness and Injury in Humans

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Table 61.4  Pesticide Exposures by Pesticide Functional and Chemical Class and Patient Age, 2006–2007 6–19 years

6 years Pesticide type

19 years

Total

No.

%

No.

%

No.

%

Hypochlorite

22,013

8524

38.7

2454

11.1

8683

39.4

Phenol

8579

5623

65.5

948

11.1

1569

18.3

Pine oil

5029

2926

58.2

441

8.8

1338

26.6

Other

11,841

7567

63.9

1151

9.7

2494

21.1

Total

47,462

24,640

51.9

4994

10.5

14,084

29.7

Sulfuryl fluoride

424

64

15.1

73

17.2

238

56.1

Other

340

44

12.9

28

8.2

201

59.1

Total

764

108

14.1

101

13.2

439

57.5

Fungicides

2260

490

21.7

167

7.4

1211

53.6

Disinfectants

Fumigants

Herbicides Chlorphenoxy compounds

4039

1173

29

353

8.7

2078

51.4

Glyphosate

8319

2490

29.9

598

7.2

4601

55.3

Other

4850

1168

24.1

406

8.4

2681

55.3

Total

17,208

4831

28.1

1357

7.9

9360

54.4

Boric acid/borates

8394

7035

83.8

310

3.7

830

9.9

Carbamates

4898

1900

38.8

446

9.1

1977

40.4

Chlorinated hydrocarbons

1042

418

40.1

132

12.7

382

36.7

Organophosphates

7680

2105

27.4

820

10.7

3840

50

Pyrethrins

10,235

3861

37.7

1221

11.9

4136

40.4

Others and combinations

70,577

23,962

34

6512

9.2

32,299

45.8

Total

102,826

39,281

38.2

9441

9.2

43,464

42.3

Insect, DEET

15,076

9861

65.4

2651

17.6

1997

13.2

Moth

7706

4764

61.8

486

6.3

1729

22.4

Other

4452

3056

68.6

475

10.7

739

16.6

Total

27,234

17,681

64.9

3612

13.3

4465

16.4

Anticoagulant

25,140

22,106

87.9

723

2.9

1840

7.3

Other

5734

3812

66.5

363

6.3

1128

19.7

Total

30,874

25,918

83.9

1086

3.5

2968

9.6

228,628

112,778

49.3

20,758

9.1

75,991

33.2

Insecticides

Repellants

Rodenticides

Total

Hayes’ Handbook of Pesticide Toxicology

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Table 61.5  Top Five Pesticide Category Exposures by Age Group, as a Percentage of All Reported Pesticide Exposures, 2006–2007 Ranking

6–19 years

6 years

19 years

Substance

%

Substance

%

Substance

%

1

Anticoagulant rodenticides

9.006

Pyrethrin/pyrethroid insecticides

2.061

Pyrethrin/pyrethroid insecticides

9.960

2

Pyrethrin/pyrethroid insecticides

6.187

DEET insect repellants

1.080

Other insecticidesa

4.205

3

Other insecticidesa

4.839

Other insecticidesa

1.016

Hypochlorite disinfectants

3.537

4

DEET insect repellants

4.017

Hypochlorite disinfectants

1.000

Glyphosate herbicides

1.874

5

Other disinfectants

b

4.008

Other disinfectants

b

0.644

Other herbicides

c

1.092

a

“Other insecticides” include insecticides other than boric acid/borates, carbamates, chlorinated hydrocarbons, organophosphates, and pyrethrins/ pyrethoids. b “Other disinfectants” include disinfectants other than hypochlorite, phenol, and pine oil. c “Other herbicides” include herbicides other than chlorophenoxy compounds and glyphosate.

Table 61.6 presents information on cases that were followed and determined to have symptoms related to their exposure. (Note that only 44.7% of cases receive follow-up to determine a final outcome.) Exposures that are expected to have no effect or minor effects, in the judgment of the SPI, often receive no follow-up and account for 46.8% of all exposures reported to the PCC. An additional 3.6% of all exposures are lost to follow-up. The remaining 2.5% of exposures involve individuals with health effects judged to be unrelated to their exposure. Among cases that received follow-up to determine medical outcome, SPIs determined that 20.7% of all pesticide exposure cases developed exposure-related symptoms. The percentage of exposures with symptoms varied with pesticide class. Symptoms developed in only 1.5% for anticoagulant rodenticide exposures, 3.8% for boric acid insecticides, up to 32.9% for other fumigants, and 32.8% for DEET-containing repellants. The low percentage of symptomatic anticoagulant rodenticide exposures is likely related to the high percentage of exposures that occur among children younger than 6 years who ingest small “taste” amounts. Parents of these children contact a poison center before symptoms have developed. The overwhelming majority of exposures are to anticoagulant rodenticides, which have very low toxicity to humans unless ingested in high quantity or repeatedly over a short period of time. Major outcomes combined with fatal outcomes were compared to all symptomatic pesticide exposures (Table 61.7). Exposure to rodenticides of all types ranked highest in severity, with 6.3% of cases in the major combined with fatal outcomes. This severe outcome is undoubtedly the result of intentional self-harm use of these products by teens and adults. Tied for first place was exposure to fumigants other than sulfuryl fluoride, also at 6.3%.

Organophosphate insecticides at 3.5% closely followed by carbamate insecticides at 3.4% ranked second and third, respectively, for severe outcomes. Some of the more toxic organophosphate and chlorinated hydrocarbon insecticides have been removed from the market. This has no doubt contributed to the decreased incidence in severe pesticide poisoning. Overall, only 0.8% of all pesticide exposure cases involved a severe outcome, but that still represents 365 patient cases during a 2-year period. (g) Discussion The NPDS data system is an important source of surveillance data of pesticide-related illness and injury. The system has several strengths. Among these are the large number of cases that are reported with approximately half (44.7%) of the reported exposures receiving follow-up. Follow-up with one call-back occurs in 50% of these cases, and multiple (range, 2–135) follow-up calls are made in the other 50% of cases. With this follow-up, the severity of the medical outcome is determined (Bronstein et al., 2007, 2008). The NPDS system also has limitations. PCCs do not collect all poison cases because reporting of poison cases is not mandatory. Many poisoning cases seen in emergency rooms, clinics, or by private physicians do not result in calls to a PCC. NPDS data represent approximately 57% of poisonings according to estimates by the Institute of Medicine. As a result, any study using PCCs as a source for poisoning cases can only be judged representative of the universe of exposures reported to PCCs and not the entire universe of all poison exposures. PCC data are a simple form of a case series and therefore not appropriate for complicated statistical analysis. However, given that the entire population of the United States and its territories is served by PCCs participating in NPDS and the large number

Chapter | 61  Surveillance of Pesticide-Related Illness and Injury in Humans

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Table 61.6  Pesticide Category and Patient Outcomes by Severity Pesticide category

Total exposures

Treated in an HCF

Minor

Moderate

Major

Death

Hypochlorite

22,013

4695

5538

915

27

1

Phenol

8579

882

1789

158

8

0

Pine oil

5029

829

1176

100

8

0

Other

11,841

1444

2548

216

13

1

Total

47,462

7850

11,051

1389

56

2

Sulfuryl fluoride

424

70

53

9

1

0

Other

340

153

75

30

5

2

Total

764

223

128

39

6

2

Fungicides

2260

445

437

76

2

0

Chlorphenoxy compounds

4039

708

809

134

5

0

Glyphosate

8319

1381

2172

179

12

6

Other

4850

1030

1020

226

13

3

Total

17,208

3119

4001

539

30

9

Boric acid/borates

8394

521

291

29

1

0

Carbamates

4898

958

633

162

25

3

Chlorinated hydrocarbons

1042

306

155

39

4

0

Organophosphates

7680

1967

1634

361

62

11

Pyrethrins

10,235

1588

1995

351

5

1

Others and combinations

70,577

10,898

14,117

2154

78

3

Total

102,826

16,238

18,825

3096

175

18

Insect, DEET

15,076

1280

4667

269

9

0

Moth

7706

1383

546

87

5

0

Other

4452

296

956

54

2

1

Total

27,234

2959

6169

410

16

1

Anticoagulant

25,140

7140

246

118

23

1

Other

5734

1855

243

110

21

3

Total

30,874

8995

489

228

44

4

228,628

42,262

41,099

5782

329

36

Disinfectants

Fumigants

Herbicides

Insecticides

Repellants

Rodenticides

Total

Hayes’ Handbook of Pesticide Toxicology

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Table 61.7  Comparison of Patients with Severe Symptoms (Major Outcome Combined with Deaths) to All Symptomatic Patients, 2006–2007 Pesticide category

Total exposures

Total symptomatic cases

Total cases with major severity or death

% of cases with major severity or death

Hypochlorite

22,013

6481

28

0.4

Phenol

8579

1955

8

0.4

Pine oil

5029

1284

8

0.6

Other

11,841

2778

14

0.5

Total

47,462

12,498

58

0.5

Sulfuryl fluoride

424

63

1

1.6

Other

340

112

7

6.3

Total

764

175

8

4.6

Fungicides

2260

515

2

0.4

Chlorphenoxy compounds

4039

948

5

0.5

Glyphosate

8319

2369

18

0.8

Other

4850

1262

16

1.3

Total

17,208

4579

39

0.9

Boric acid/borates

8394

321

1

0.3

Carbamates

4898

823

28

3.4

Chlorinated hydrocarbons

1042

198

4

2

Organophosphates

7680

2068

73

3.5

Pyrethrins

10,235

2352

6

0.3

Others and combinations

70,577

16,352

81

0.5

Total

102,826

22,114

193

0.9

Insect, DEET

15,076

4945

9

0.2

Moth

7706

638

5

0.8

Other

4452

1013

3

0.3

Total

27,234

6596

17

0.3

Anticoagulant

25,140

388

24

6.2

Other

5734

377

24

6.4

Total

30,874

765

48

6.3

228,628

47,246

365

0.8

Disinfectants

Fumigants

Herbicides

Insecticides

Repellants

Rodenticides

Total

Chapter | 61  Surveillance of Pesticide-Related Illness and Injury in Humans

of poison exposures, PCC data are likely to be helpful for identifying exposure situations that should be targeted for risk mitigation (Institute of Medicine, 2004). Misclassification may occur when symptoms are selfreported over the phone and are not confirmed by a physician or by laboratory tests. PCC data reflect only what is reported by the lay public or health professional when the PCC is called. Information may be incomplete because it may include what the caller thought was important but may leave out important details. Although 15.5% of calls to PCCs are made by health care professionals, the majority are calls made by the victim or their relative. The PCC SPI must rely on his or her experience and judgment to determine which cases have symptoms consistent with the toxic substance, dose, and timing of the exposure. When health effects are judged to be unrelated to the exposure, evidence should support this determination and should be documented in the PCC case record. Although some misclassification can be expected to occur, it is assumed to be nondifferential across pesticides. That is, there is no reason to believe that SPIs are likely to misclassify one pesticide more or less than another. Although reports in NPDS are labeled “poison exposures,” many of the cases in the database never develop health effects as a result of the exposure. Several potential explanations exist for the lack of health effects in these cases, include the following: Advice provided by the PCC led to prompt treatment (e.g., removing the toxic substance from the skin, rapidly removing the patient from a spray area, and irrigating the eyes thoroughly), the exposure agent was relatively nontoxic, the exposure dose was not great enough to produce toxicity, and the exposure was suspected but actually never occurred (cases coded as “confirmed nonexposures” in which there is sufficient evidence that an exposure never occurred are removed from NPDS). Tracking asymptomatic exposure cases can be useful because asymptomatic cases can be an indicator of the number of pesticide exposures usually occurring at home that are not reported elsewhere. Limitations involving some severity measures used in this review may be present. When using data obtained from the public in circumstances they consider to be an emergency, inaccuracy is inevitable. A misperception concerning the danger or risk of certain types of products rather than the actual risks may exist. For example, both the public and some health care professionals may perceive ingestion of “rat poison” as more dangerous than ingestion of other types of pesticides. As a result, such cases are more likely to be seen in a health care facility even though the overwhelming majority of cases involve minor exposures to anticoagulants (e.g., just a taste) that pose little risk. Of pesticide exposures, 49.3% occur in children younger than 6 years. There is a strong possibility that parents may panic when their child ingests a substance called “rat poison,” “ant killer,” or “roach killer.” Parents may take the child to the emergency room immediately without

1323

calling for help. Decisions about which cases become hospitalized may be due to inaccurate perceptions of risk and differences in the health care professional’s experience. Obtaining medical care by patients is affected by availability and extent of health insurance coverage or workers’ compensation. With PCC advice, triage of patients to the appropriate treatment site results in significant savings in health care costs. PCCs manage 72.8% of cases at home, and this can prevent unnecessary HCF visits. One study showed that for every $1 spent by the poison center, $7 in unneeded health care is saved (Miller and Lestina, 1997). For those with pesticide exposures, Table 61.8 compares the incidence of patients presenting to an HCF with the incidence of patients who develop symptoms. Overall, 18.5% of pesticide-exposed patients present to a HCF and 20.7% develop symptoms. Exposures to rodenticides result in 29.1% of patients presenting to an HCF but only 2.5% actually develop symptoms, contrasting with 8.5% of patients who present to a HCF with exposures to DEET repellants, while 32.8% develop symptoms. For pesticideexposure cases called to PCCs, more patients are kept at home than are seen in an HCF. (h) Syndromic Surveillance The initial purpose of NPDS was poison center data collection. However, it soon became clear that NPDS, with its almost real-time data collection (PCC cases are uploaded a mean of every 14 min), was important as a public health response network. The CDC and AAPCC are working together to develop a true national poison center data system. Incoming data are continually monitored for any anomalous signal detection. If detected, the signal alerts a member of the AAPCC NPDS surveillance team. If the detected signal is significant, the surveillance team member can contact local PCCs. NPDS can also generate notifications or alerts on items of public health interest, such as adverse reactions, product recalls, or contaminated food. Notifications and alerts on public health issues are also sent to the National Center for Environmental Health at the CDC (Bronstein et al., 2007, 2008). From September 2006 through 2008, more than 100,000 anomalies were detected. In 2007, there were 352 surveillance definitions continuously running to monitor NPDS data by case anomaly definition and clinical effects (e.g., multiple victims exposed and food poisoning, nerve agents). Surveillance enhancement and software improvements have allowed the addition of case drill-down using GIS (geographic information system). Although PCC data have not identified any index cases, close work with public health agencies shows promise as part of an early detection program. More extensive partnerships with governmental agencies for surveillance purposes are anticipated in the near future. However, as local poison center funding disappears, the major roadblock will be finding the funding (Bronstein et al., 2008).

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Table 61.8  Number of Patients Treated in a Health Care Facility Compared to the Number of Patients with Symptoms Pesticide category

Total exposures

Patients treated in an HCF

Symptomatic patients

No.

%

No.

%

Disinfectants Hypochlorite

22,013

4695

21.3

6481

29.4

Phenol

8579

882

10.3

1955

22.8

Pine oil

5029

829

16.5

1284

25.5

Other

11,841

1444

12.2

2778

23.5

Total

47,462

7850

16.5

12,498

26.3

Sulfuryl fluoride

424

70

16.5

63

14.9

Other

340

153

45

112

32.9

Total

764

223

29.2

175

22.9

Fungicides

2260

445

19.7

515

22.8

Chlorphenoxy compounds

4039

708

17.5

948

23.5

Glyphosate

8319

1381

16.6

2369

28.5

Other

4850

1030

21.2

1262

26

Total

17,208

3119

18.1

4579

26.6

Boric acid/borates

8394

521

6.2

321

3.8

Carbamates

4898

958

19.6

823

16.8

Chlorinated hydrocarbons

1042

306

29.4

198

19

Organophosphates

7680

1967

25.6

2068

26.9

Pyrethrins

10,235

1588

15.5

2352

23

Others and combinations

70,577

10,898

15.4

16,352

23.2

Total

102,826

16,238

15.8

22,114

21.5

Insect, DEET

15,076

1280

8.5

4945

32.8

Moth

7706

1383

17.9

638

8.3

Other

4452

296

6.6

1013

22.8

Total

27,234

2959

10.9

6596

24.2

Anticoagulant

25,140

7140

28.4

388

1.5

Other

5734

1855

32.4

377

6.6

Total

30,874

8995

29.1

765

2.5

228,628

42,262

18.5

47,246

20.7

Fumigants

Herbicides

Insecticides

Repellants

Rodenticides

Total

Chapter | 61  Surveillance of Pesticide-Related Illness and Injury in Humans

61.2.2  State-based Surveillance Systems (a) Description Thirty states in the United States require some form of physician, laboratory, or hospital reporting of pesticiderelated illness. These states are listed in Table 61.9 along with information on the specifics of the reporting rule. Only 12 states (Arizona, California, Florida, Iowa, Louisiana, Michigan, New Mexico, New York, North Carolina, Oregon, Texas, and Washington) routinely conduct more comprehensive case investigation and surveillance activities. In response to both public concern and the successes achieved by the 12 states currently conducting surveillance, other states are considering initiating pesticide poisoning surveillance activities. Before 1998, the state-based pesticide poisoning surveillance systems used a variety of methods for collecting and categorizing data that did not permit the routine pooling and analysis of multistate data. Recognizing the importance of systematic surveillance efforts, NIOSH, with

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funding assistance from the U.S. EPA, led efforts to standardize pesticide poisoning surveillance. A collaboration involving experts from federal agencies (NIOSH, U.S. EPA, and National Center for Environmental Health), nonfederal agencies (Council of State and Territorial Epidemiologists and Association of Occupational and Environmental Clinics), and state health departments or other state designees developed a standardized set of variables for pesticiderelated illness and injury surveillance. This standardized set of variables, including a standardized case definition (described elsewhere in this chapter), was finalized in 1998 and is used by the 12 states mentioned previously. The large number of pesticide products on the market and difficulties in obtaining case reports make the pooling of all available data particularly desirable. Having standardized variables and a standardized case definition facilitates the aggregation of these data. The aggregated data have proven useful to regulatory agencies, public health policymakers, researchers, worker education programs, the public, and the medical community.

Table 61.9  Pesticide-Related Illness Mandated Reporting Requirements and Entities by Statea State

Pesticide reporting requirementb

Entities mandated to report

Physician

Hospital

Laboratory

Poison control center

Other health care professional

X

X

Alaska

ANY OCC DZ

X

Arizona

ANY PEST

X

Arkansas

ANY PEST

X

California

ANY PEST

X

Florida

ANY PEST

X

Hawaii

ANY PEST

X

Illinois

ANY OCC DZ

X

X

X

Iowa

ANY PEST

X

X

X

Louisiana

ANY PEST

X

X

X

Maine

ANY OCC DZ

X

X

Maryland

ANY PEST

X

X

Massachusetts

OCC PEST

X

X

Michigan

ANY OCC DZ

X

X

X

Mississippi

ANY PEST

X

X

X

X

Missouri

ANY PEST

X

X

X

X

Nebraska

ANY PEST

X

X

X

X X X

X

X

X

X

X

X

(Continued)

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1326

Table 61.9  (Continued) State

Pesticide reporting requirementb

Entities mandated to report Physician

Hospital

Laboratory

New Hampshire

ANY OCC DZ

X

New Jersey

ANY PEST

X

X

New Mexico

OCC PEST

X

X

X

New York

ANY PEST

X

X

X

North Carolina

ANY PEST

X

Ohio

OCC PEST

X

Oregon

ANY PEST

X

X

X

South Carolina

ANY PEST

X

X

X

Texas

OCC PEST

X

X

X

Utah

ANY TOXIN

X

X

X

Virginia

ANY PEST

X

X

X

Washington

ANY PEST

X

Wisconsin

OCC PEST

X

X

Wyoming

ANY TOXIN

X

X

Poison control center

Other health care professional

X

X

Sources of data: Calvert et al. (2001); Internet search; personal communication with Erin Simms, Council of State and Territorial Epidemiologists; and e-mails and calls to selected state agencies to clarify inconsistencies. a This table does not include states with only voluntary pesticide reporting requirements (Idaho, Illinois, North Dakota, South Dakota, and Vermont). b ANY PEST, reporting of any pesticide-related illness (whether occupational or nonoccupational) is mandated; OCC PEST, only reporting of occupational pesticide-related illness is mandated (there are no requirements for reporting poisoning from nonoccupational toxic exposures); ANY OCC DZ, reporting of any occupational disease is mandated (there are no specific requirements for pesticide-related illness reporting, nor are there requirements for reporting poisoning from nonoccupational toxic exposures); ANY TOXIN, reporting of any poisoning from toxic exposures is mandated (there are no specific requirements for pesticide-related illness reporting); OCC TOXIN, reporting of any poisoning from occupational toxic exposures is mandated (there are no specific requirements for pesticide-related illness reporting, nor are there requirements for reporting poisoning from nonoccupational toxic exposures).

This section briefly describes the surveillance systems in the 12 states mentioned previously. Collectively, these states are referred to as the Sentinel Event Notification System for Occupational Risk (SENSOR)-Pesticides program. The SENSOR-Pesticides states have much in common and are not described separately in detail. However, there are two pesticide poisoning surveillance programs in California. One is maintained by the California Department of Pesticide Regulation (DPR) and the other by the California Department of Public Health (CDPH). The surveillance system maintained by DPR is described in a separate section because it uses a slightly different case definition and variables. The DPR system is also the largest system (both in poisoning cases annually identified and in staffing levels) and has been in existence longer than any other surveillance system with the possible

exception of Washington State. The surveillance system maintained by CDPH participates in the SENSORPesticides program and is similar to the other surveillance systems described in this section. However, unlike the other SENSOR-Pesticides systems, the CDPH surveillance system tracks occupational pesticide-related illness and injury only. The state-based surveillance programs use multiple sources for case ascertainment (see Data Source), and active case follow-up is performed either directly by the surveillance system or conducted by partner state agencies (e.g., the state agriculture department). Several of these state systems originally included a system of sentinel health care professionals who were contacted on a regular basis. This approach was labor-intensive, did not yield many cases, and was discontinued (Schnitzer and Shannon, 1999).

Chapter | 61  Surveillance of Pesticide-Related Illness and Injury in Humans

Once a report is received, participating SENSORPesticides states review the information available in the report to determine whether the subject was symptomatic and whether the involved chemical was a pesticide. If so, attempts are made to interview the poisoned subject or his or her proxy to obtain details on the poisoning event. If the subject sought health care, a request is made for relevant medical records. The program may also interview other knowledgeable parties, including the supervisor of the exposed person, the applicator, and witnesses. The information collected by the state agencies in a standardized manner includes date of illness, information on the ill individual (gender, race, age, signs, and symptoms, industry, and occupation), whether the illness occurred as a result of workplace exposures, identification of the pesticide(s) that produced the illness, activity of the individual when exposed, type of exposure (e.g., drift, direct spray, indoor air exposure, or exposure to a spill or leaking container), biological monitoring information (i.e., cholinesterase testing and results and whether other biological testing was performed), factors that contributed to exposure (e.g., lack of notification of pesticide application, early re-entry into a treated area, and equipment failure), and use of personal protective equipment (PPE). All states that participate in the SENSOR-Pesticides program receive funding support from their state. State funding can come from one or more of the following sources: general funds, licensing fees, or pesticide product registration fees. In addition, seven SENSOR-Pesticides states are partially funded by NIOSH and the U.S. EPA through the SENSOR-Pesticides program. Besides collecting information on each suspected case of pesticide poisoning, these SENSOR-Pesticides surveillance systems perform in-depth investigations for cases that meet priority criteria, identify trends and emerging pesticide problems, prepare reports of investigations and other program activities, and develop interventions aimed at particular industries or hazards. The state-based SENSOR-Pesticides surveillance programs do not have regulatory authority. In addition to providing funding to some SENSORPesticides states, NIOSH plays several other roles in the SENSOR-Pesticides program. NIOSH provides advice and guidance to states in developing and maintaining their surveillance system. To foster communication with participating states, NIOSH maintains a listserv and website for the program. NIOSH participates in the SENSOR-Pesticides coding committee, which also includes representatives from some SENSOR-Pesticides states. The purpose of this committee is to identify and recommend changes to the standardized variables used by the SENSOR-Pesticides program. NIOSH creates and maintains a national aggregated database of pesticide poisoning cases identified by states participating in the SENOR-Pesticides program. NIOSH also prepares and publishes manuscripts in Mortality and Morbidity Weekly Report and peer-reviewed

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journals. SENSOR-Pesticides state partners rely on NIOSH to produce these papers because the states often do not have the time or expertise for such efforts. (b) Case Definition Beginning with cases identified and exposed to pesticides in 1998 or later, the SENSOR-Pesticides program began using the National Public Health Surveillance System case definition and classification scheme to evaluate reports. This case definition is described later (CDC, 2001a). Briefly, information in three areas is required: pesticide exposure, health effects, and toxicological evidence supporting an association between exposure and effect. A case of pesticide-related illness or injury is classified into one of the following categories: definite, probable, possible, or suspicious. The specific classification category applied to a given case depends on the certainty of exposure, whether health effects consisted of signs observed by a health care professional versus symptoms reported by the poisoned subject, and the extent to which the health effects were consistent with the known toxicology of the pesticide product. Illnesses associated with intentional (e.g., suicidal and malicious intent) exposures were excluded from the findings presented here. Reports can also be classified into one of four other categories: unlikely case (untenable exposure– health effect relationship), insufficient information, unrelated (illness determined to be caused by a condition other than pesticide exposure), and asymptomatic. Reports classified into one of these four categories are not considered cases of pesticide-related illness or injury. Illness severity is also assigned to all cases using standardized criteria that are based on signs and symptoms, medical care received, and lost time from work (CDC, 2001b). Low severity illness/injury consists of illnesses and injuries that generally resolve without treatment and where minimal time (3 days) is lost from work. Such cases typically manifest as eye, skin, and/or upper respiratory irritation. Moderate severity illness/injury consists of non-life-threatening health effects that are generally systemic and require medical treatment. No residual disability is detected, and time lost from work is less than 6 days. High severity illness/injury consists of life-threatening health effects that usually require hospitalization, involve substantial time lost from work (5 days), and may result in permanent impairment or disability. Death pertains to fatalities resulting from exposure to one or more pesticides. (c) Data Source All of these surveillance systems require reporting of pesticide-related illness and injury cases from physicians (see Table 61.9). However, relatively few cases are directly reported by physicians. When SENSOR was originally conceived, cases were to be reported to state

1328

health departments by a targeted group of health care professionals, or sentinel providers (Baker, 1989). However, this concept of SENSOR was abandoned because of the time and expense needed to periodically remind clinicians to report cases (Levy et al., 1992; Schnitzer and Shannon, 1999). The principal source of case reports varies across states but is generally one of the following: workers’ compensation claims, poison control centers, and state agencies with jurisdiction over pesticide use (e.g., state agricultural departments). States also routinely review other data sources to identify additional potential cases and to evaluate the completeness of reporting. Other sources of case reports include emergency medical services, medical laboratories, hospital emergency rooms, hospital discharge data, clinics, Migrant Legal Aid, selected community contacts, state structural pest control boards, and death certificates. Most states accept reports from sources other than those required to report through regulation. Some states also accept self-reports as a trigger for investigation. Both Oregon and Washington maintain interagency boards that are required to coordinate the investigation of reported adverse impacts from pesticides, review incidents, and develop strategies to prevent exposures. The interagency board in Oregon is called the Pesticide Analytical and Response Center (PARC), and the Washington board is called the Pesticide Incident Reporting and Tracking Review Panel (PIRT) (Barnett and Calvert, 2005). Both interagency boards are composed of representatives from agencies with jurisdiction over pesticides, health, and the environment. In addition, these interagency boards include a state poison control center representative and an appointed general public member. PIRT also includes a practicing toxicologist and representation from the state universities. The state universities serve as consultants to the PARC board but are not members.

(d) Target Population These systems strive to capture any pesticide-related acute illness or injury occurring in the state population. The systems capture illness and injuries resulting from both occupational and nonoccupational exposures, with the exception of the CDPH. In contrast to the DPR, which captures both occupational and nonoccupational cases, CDPH captures occupational cases only. (e) Period of Time of Data Collection As explained previously, the SENSOR-Pesticides standardized case definition and standardized variables were finalized in 1998. This permitted the creation of a national aggregated data set beginning in that year. The national aggregated data set includes data from states participating in the SENSOR-Pesticides program (Figure 61.3).

Hayes’ Handbook of Pesticide Toxicology

Figure 61.3  States that participated in the SENSOR-Pesticides program in 2009 (in white, n  12).

However, in some states, the commencement of acute pesticide-related illness and injury surveillance occurred before 1998. Of the SENSOR-Pesticides states, Washington has the oldest system, having been established in the early 1970s (note that the DPR also had a surveillance system and a reporting law established in the early 1970s). However, Washington did not have an acute pesticide poisoning reporting requirement until 1989. Washington surveillance data are considered complete beginning with calendar year 1991. Of the SENSOR-Pesticides states (excluding California), Texas has the oldest acute pesticide poisoning reporting rule, which was enacted in 1986, and surveillance data are considered complete beginning with calendar year 1988. Oregon has required health care providers to report pesticide poisoning since 1987, and surveillance data are considered complete beginning with calendar year 1988. In Arizona, although the reporting rule went into effect in 1987, surveillance data are considered complete beginning with the calendar year 1992. Likewise, in Florida, the reporting rule went into effect in 1987; however, reporting had been limited until the more comprehensive surveillance system was initiated in 1997 and surveillance data are considered complete beginning with calendar year 1998. In New York, acute pesticide poisoning has been a reportable condition since August 1990, and surveillance data are considered complete beginning with calendar year 1991. Although New Mexico has had a reporting requirement since 1993, its surveillance system was not fully developed until 2004, and surveillance data are considered complete beginning in 2005. Louisiana and Michigan have complete data beginning in 2001, and Iowa and North Carolina have complete data beginning in 2006. In each of the SENSOR-Pesticides states, complete data for a calendar year are generally available 5–12 months after the end of the calendar year. (f) Periodicity of Reports Printed reports are generally published annually.

Chapter | 61  Surveillance of Pesticide-Related Illness and Injury in Humans

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Table 61.10  Acute Pesticide-Related Illness by Functional Class and Severity, SENSOR-Pesticides, 1998–2005 Pesticide functional class

Severity category

Total (%)

Low

Moderate

High

Death

Insecticides

3826

895

66

12

4799 (57)

Herbicides

935

156

6

0

1097 (13)

Fungicides

180

41

5

0

226 (3)

Fumigants

223

99

4

3

329 (4)

Disinfectants

499

80

7

1

587 (7)

Insecticides  fungicides

222

56

3

0

281(3)

Othera

507

74

11

4

596 (7)

Multipleb

417

114

5

1

537 (6)

6809 (81%)

1515 (18%)

107 (1.3%)

21 (0.3%)

8452

Total a

This category includes rodenticides, plant growth regulators, insect growth regulators, wood treatment products, preservatives, and insect repellants. Pesticide product was classified into more than one functional class.

b

1200

Rate per million full time equivalents

(g) Findings Between 1998 and 2005, a total 8452 cases of acute pesticide-related illness or injury were identified by the SENSOR-Pesticides program. A total of 4512 cases were work-related and 3940 were nonoccupational pesticiderelated illnesses or injuries. These cases fell into the following classification categories: definite, 826 (10%); probable, 1793 (21%); possible, 5330 (63%); and suspicious, 503 (6%). The mean age was 36 years (range, 1 month–99 years). Males accounted for 53% of the cases (62% of occupational cases and 44% of nonoccupational cases). Most illnesses were of low severity (80%) (Table 61.10). A total of 21 unintentional pesticide-related deaths were identified between 1998 and 2005. These included 3 agricultural workers, 4 nonagricultural workers, and 14 nonoccupational cases. The average annual occupational pesticide poisoning incidence rate for 1998–2005 was 10.8 per million fulltime equivalents (FTEs). FTE estimates, which are estimates of the number of full-time workers in the United States, were derived from the Current Population Survey conducted between 1998 and 2005 (Bureau of Labor Statistics, 2007). Figure 61.4 provides average annual incidence rates by state. The average incidence rates for all workers ranged from 2.2 per million FTEs in Arizona to 37 per million FTEs in Washington. The incidence rate per million workers also decreased from 13.1 in 1998 to 10.5 in 2005 (p value for trend by Poisson regression test 0.001). A total of 1627 cases were employed in agriculture (39% of occupational cases where industry information was available) (Figure 61.5). The average annual agricultural

1000 800 600 400 200 0 AZ

CA

FL

LA

MI

NM

NY

OR

TX

WA

State Agriculture

Total

Figure 61.4  Average annual incidence rate of occupational pesticide poisoning by state, SENSOR-Pesticides, 1998–2005.

pesticide poisoning incidence rate for 1998–2005 was 239 per million FTEs. The incidence rate per million also decreased from 384 in 1998 to 318 in 2005 (p  0.001). The annual agricultural pesticide poisoning incidence rate was as low as 174 per million FTEs in 2004. For the years 1998–2005, New York had the lowest average annual agricultural pesticide poisoning incidence rate (45 per million FTEs) and Washington had the highest (1057 per million FTEs) (see Figure 61.4). A total of 2885 (34%) cases were employed in nonagricultural industries (61% of occupational cases where industry information was available). Most of these cases were employed in the services sector (see Figure 61.5).

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1330

Figure 61.5  Distribution of occupational cases by industry, 1998–2005, SENSOR-Pesticides (n  4131).

Table 61.11  Acute Pesticide-Related Illness by Functional Class and Occupational Status of Victim, SENSOR-Pesticides, 1998–2005 Pesticide functional class

Occupational status of victim

Total (%)

Agricultural workers, n (%)a

Nonagricultural workers, n (%)a

Nonoccupational cases, n (%)

Insecticides

635 (39)

1604 (56)

2558 (65)

4799 (57)

Herbicides

246 (15)

402 (14)

450 (11)

1097 (13)

Fungicides

125 (8)

53 (2)

48 (1)

226 (3)

Fumigants

146 (9)

139 (5)

44 (1)

329 (4)

Disinfectants

56 (3)

290 (10)

241 (6)

587 (7)

Insecticides  fungicides

175 (11)

57 (2)

49 (1)

281(3)

80 (5)

190 (7)

327 (8)

596 (7)

164 (10)

150 (5)

223 (6)

537 (6)

1627 (19)

2885 (34)

3940 (47)

8452

Other

b

Multiple Total

c

a

Agricultural workers were defined as workers whose industry was coded as agricultural production, excluding livestock [1990 Census Industry Code (CIC)  010]; agricultural production, including livestock (1990 CIC  011); and agricultural services (1990 CIC  030; 2002 CIC  0290) (U.S. Bureau of the Census, 1992). Nonagricultural workers included workers employed in all other industries. b This category includes rodenticides, plant growth regulators, insect growth regulators, wood treatment products, preservatives, and insect repellants. c Pesticide product was classified into more than one functional class.

Information on the pesticides responsible for acute occupational pesticide-related illness is provided in Tables 61.10, 61.11, and 61.12. Overall, insecticides were responsible for 57% (n  4799) of the illnesses. Among the insecticides, organophosphates (n  1760), pyrethroids (n  1305), pyrethrins (n  335), carbamates (n  276), mixtures involving cholinesterase inhibitors (n  200), and mixtures

of pyrethrins and pyrethroids (n  162) were most commonly responsible. Table 61.11 provides the distribution of pesticide functional classes by occupational status, including agricultural workers, nonagricultural workers, and nonoccupational cases. In all three categories, insecticides were responsible for the largest proportion of illnesses, followed by

Chapter | 61  Surveillance of Pesticide-Related Illness and Injury in Humans

1331

Table 61.12  Top 15 Pesticide Active Ingredients Most Commonly Involved in Illness by Occupational Status, SENSOR-Pesticides, 1998–2005 Rank

Agricultural workers (N  1627)a

No.

Nonagricultural workers (N  2885)a

No.

Nonoccupational cases (N  3940)

No.

1

Sulfur

186

Pyrethrins

326

Pyrethrins

525

2

Chlorpyrifos

121

Chlorpyrifos

170

Malathion

417

3

Glyphosate

90

Malathion

157

Permethrin

300

4

Cyfluthrin

81

Diazinon

131

Chlorpyrifos

259

5

Glyphosate

79

Resmethrin

130

Diazinon

196

6

Malathion

69

Permethrin

122

Cypermethrin

153

7

Methyl bromide

68

Glyphosate

113

DEET

142

8

Mancozeb

59

2,4-D

111

Tetramethrin

139

9

Carbofuran

58

Acephate

98

Glyphosate

120

10

Propargite

57

Silica gel

89

Propoxur

119

11

Imidacloprid

57

Cypermethrin

85

Phenothrin

111

12

Abamectin

56

Quaternary ammonium compounds

77

Cyfluthrin

90

13

Mepiquit chloride

52

Abamectin

71

d-trans-Allethrin

82

14

Paraquat dichloride

49

d-trans-Allethrin

64

Sodium hypochlorite

79

15

Carbaryl

47

Sodium hypochlorite

62

2,4-D

78

a Agricultural workers were defined as workers whose industry was coded as agricultural production, excluding livestock [1990 Census Industry Code (CIC)  010]; agricultural production, including livestock (1990 CIC  011); and agricultural services (1990 CIC  030; 2002 CIC  0290) (U.S. Bureau of the Census, 1992). Nonagricultural workers included workers employed in all other industries.

herbicides. Table 61.12 provides the active ingredients most commonly associated with illness, stratified by occupational status. (h) Discussion Among the strengths of these state systems is their reliance on a broad number of sources for case ascertainment. For example, the development of close ties with regional poison control centers has served to provide more complete reporting, particularly for illnesses from nonoccupational exposures. In addition, workers’ compensation systems can be an important source of occupational cases as documented in California and Washington. Other states should consider improving access to workers’ compensation data because these data are an important source of cases and can be used to periodically evaluate the completeness of the surveillance system. Another strength of state-based surveillance systems is their access to personal identifiers. By knowing the identity of a case and the location of the exposure, prompt appropriate follow-up and intervention can be instituted. For example, in the case of an occupational pesticide-related illness, identification of the responsible workplace can result in an

investigation to identify other workers with illness and to precisely target appropriate prevention programs. This is in contrast to national surveillance systems, which provide only anonymous data without personal identifiers. Many state systems have found that maintaining physician (or any health care provider-based) reporting is resource intensive (Schnitzer and Shannon, 1999). When some states attempted to promote physician reporting through outreach activities, they found that case reporting increased but only as long as the outreach activities persisted. For physician reporting to be successful, the health care professional must be able to recognize pesticide-related illness and must comply with reporting requirements in a timely manner. Considering that pesticide-related illness is relatively rare and that health care professionals may not be trained in its recognition, the expectation that the preceding steps occur is likely over­ optimistic. These issues are explored in more depth later. Some states require laboratories to report when test results yield evidence for pesticide-related illness. The cholinesterase test is probably the most common laboratory test for recognizing pesticide-related illness; however, it is only useful for organophosphate and carbamate pesticides. An additional limitation of cholinesterase reports

1332

is that only a minority of them are associated with known organophosphate or carbamate exposures. For example, in New York, where state law mandates that clinical laboratories report abnormally depressed cholinesterase levels, of 198 such laboratory reports received in 1995 and 1996, only 62 (31%) had known pesticide exposures (New York State Department of Health, 1997). The remaining cases were often tested as part of a presurgical evaluation and had either congenitally low levels or illnesses associated with low cholinesterase levels (e.g., liver disease, malnutrition, acute infections, and pernicious anemia; Vorhaus and Kark, 1953). Before deciding to adopt laboratory reporting for cholinesterase levels, consideration must be made for the resources needed to follow up on abnormal results. The data provided by these state surveillance systems have proved useful. In addition to assessing magnitude and identifying risk factors, the SENSOR-Pesticides program has identified many emerging pesticide problems, usually in collaboration with DPR (Alarcon et al., 2005; Calvert et al., 2003, 2006, 2007a,b, 2008; Calvert and Higgins, 2009; CDC, 1999a,b,c, 2000, 2003, 2004, 2008a,b; Das et al., 2001). These reports led to targeted efforts to prevent the recurrence of the identified pesticide problem. For example, after illnesses were found to be associated with the pesticides used in total release foggers (TRFs) (CDC, 2008a), New York began the process of classifying TRFs as restricted-use products (New York State Department of Environmental Conservation, 2008). SENSOR-Pesticides also recommended TRF package redesign, label modification, and public awareness campaigns. Another emerging pesticide problem that was detected involved illnesses associated with pesticide exposures at schools (Alarcon et al., 2005). The publication of this report provided additional impetus for passage of legislation in several states that requires schools and school districts to reduce pesticide usage by implementing integrated pest management programs. Another report documented the problem with offtarget drift of pesticides (CDC, 2003), specifically off-target drift into a low-income Hispanic community where many residents lacked health insurance. After this report was published, a law was passed in California that made growers liable for the uncompensated medical care provided to those who become sick from pesticides that drift from the grower’s farm (State of California Food and Agriculture Code, Sections 12996.5, 12997.5, and 12997.7). The SENSOR-Pesticides program has at least two limitations. First, the incidence rates are likely to be underestimates due to several factors (Azaroff et al., 2002). Many individuals with pesticide-related illness are never ascertained because they neither seek medical care nor call appropriate authorities. Furthermore, because the signs and symptoms of acute pesticide-related illness are not pathognomonic, and because most health care professionals receive little instruction on this illness, many who seek medical care may not be correctly diagnosed. Even among

Hayes’ Handbook of Pesticide Toxicology

those who are correctly diagnosed, many are not reported to state surveillance systems. Second, because only 12 states participate in the program, national estimates of pesticide poisoning are not available.

61.2.3  California Department of Pesticide Regulation (a) Description A 1971 California law requires physicians to report to the local health officer any disease or condition that they know or have reason to believe to be caused by pesticide exposure. The local health officer must notify the county agricultural commissioner (CAC) immediately. The health officer has 7 days to transmit the report to three state agencies, including the DPR. DPR maintains the California Pesticide Illness Surveillance Program (PISP), which supplements physician reports by reviewing reports of occupational cases forwarded to the California Bureau of Labor Statistics (CBLS) by workers’ compensation insurers. In 2006, following developmental work funded by the U.S. EPA, DPR negotiated a contract with the California Poison Control System (CPCS). Under this contract, CPCS staff members offer to fulfill the reporting requirement on behalf of physicians who consult them. CACs investigate all identified cases (including those involving nonagricultural exposures) of people exposed to pesticides in their jurisdictions. DPR provides instructions, training, and technical support for investigations. Instructions include directions for when and how to collect samples of foliage, clothing, or surface residues to document environmental exposures. As part of the technical support, DPR contracts with the California Department of Food and Agriculture’s Center for Analytical Chemistry to analyze the samples. DPR also provides guidance to CACs in investigating sensitive situations such as drift into residential areas and suicides or attempted suicides. In cases of self-harm, DPR recommends that CACs seek primarily reports filed by first responders and exercise caution about making contact with victims or their families. The CACs attempt to determine the circumstances of pesticide exposure. The investigators try to locate and interview everyone with knowledge of the event, including the affected people, the supervisors of people exposed at work, any other witnesses, and all applicators involved in the implicated application (if any) and their supervisors. Depending on what they find, investigators may track the source of the pesticide or ask the affected people to release relevant medical records. Investigators also inspect equipment and records to determine whether all laws and regulations were observed in acquiring and handling the pesticide(s) (Figure 61.6). If the investigation identifies noncompliance with laws or regulations, the CACs take action to enforce safe pesticide management.

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Figure 61.6  A twin-row, over-the-vine air blast sprayer applies pesticide to grape vines (courtesy of California EPA).

DPR scientists abstract data from the CACs’ reports and any medical records received into a computerized database. Software migration in 1998 provided the occasion to expand the database and organize its entries more logically. Cases from 1992 on have been upgraded to current standards. (b) Case Definition Physician reporting is required for “pesticide poisoning or any disease or condition caused by a pesticide.” This requirement of California Health and Safety Code Section 105200 specifically states that such consultations may not be dismissed as “first aid”; doctors must report all suspected pesticide-related cases. The law also applies to cases of suicide or attempted suicide. DPR recognizes that pesticide products are complex mixtures with various possible mechanisms of action. It is DPR policy to consider any adverse health effect that results from pesticide exposure to be a pesticide-related illness or injury. For purposes of overall classification, the primary toxic effects of the active ingredient(s) are not distinguished from effects of inert ingredients or impurities or from incidental reactions to product characteristics, such as nausea in response to odor. DPR scientists use a 5-point ordinal scale to record a qualitative assessment of the likelihood that pesticide exposure caused or contributed to the reported illness or injury. When several signs or symptoms are reported, the scientist records the relationship of the complaint most likely to result from pesticide exposure. If the case record lacks critical information, scientists do not assign a relationship. For those cases that provide enough information

to support evaluation, the relationship options consist of the following: Definite: The signs and symptoms exhibited by the affected person are such as would be expected to result from the exposure described. Both medical evidence (e.g., blood cholinesterase levels or allergy testing) and physical evidence (e.g., residue on leaf samples or contaminated clothing) support the conclusion that the illness or injury was the result of the pesticide exposure. Probable: There is close correspondence between the pattern of exposure and the illness or injury experienced. Medical and/or physical evidence may not be available. For example, although symptoms may be highly suggestive of cholinesterase inhibition, without results of cholinesterase testing, the case would have to be entered as probable rather than definite. Possible: Health effects correspond generally to the reported exposure, but evidence is not available to support a relationship. The information available may be ambiguous. Headaches, nausea, and skin rashes, for example, all can be caused by many different things, and sometimes people are uncertain about the exact sequence of exposures relative to onset of ill health. Such uncertainty will cause a case to be entered as possible. Unlikely: The exposure may be uncertain; the signs and symptoms reported are not typical of the suspected exposure, but the possibility that the victim is suffering the effects of pesticide exposure cannot be dismissed. Uncertain exposures may involve people far from the application site or who only handled tightly closed packages or thoroughly cleaned containers.

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1334

Unrelated: Evidence is available to demonstrate that the illness or injury was caused by factors other than exposure to pesticides. Sometimes, a product that initially was thought to be a pesticide turns out to be something else, such as a fertilizer or cleaner. Other times, the attending physician determines that the problem is infectious, not toxic. Asymptomatic: The subject of the investigation was exposed to one or more pesticides but suffered no illness or injury. Cholinesterase depression without symptoms falls in this category. Such cases may, however, reflect lapses from good work practice. Indirect: The illness or injury reported appears to have been caused not by pesticide exposure but by measures prescribed for avoiding pesticide exposure. People who develop heat stress through performing vigorous work in heavy protective clothing fall into this category, as do those who develop allergic reactions to rubber gloves. Probable and definite cases generally are combined in data analyses. The category of possible relationship is the most ambiguous. In practice, it indicates that the people involved are known to have had contact with pesticides shortly before becoming ill or injured, but firm evidence is not available to indicate whether or not pesticide exposure caused their illness or injury. Because the “possible” category may contain more false positives than the definite/ probable category, the PISP generally discusses possible cases separately from other categories. However, when determining the magnitude of acute pesticide-related illness and injury, cases classified as definite, probable, or possible are included in the sum, which is referred to as the number of pesticide-associated cases. Unrelated, asymptomatic, and indirect relationships all indicate an assessment that pesticide exposure had nothing to do with health effects. Cases judged unlikely to relate to pesticide exposure are typically grouped with unrelated cases in reports. (c) Data Source Until CPCS began facilitating reports, CBLS provided 60–75% of all cases investigated. In 2007, physician reports transmitted through CPCS accounted for 36% of all cases and CBLS for 33%. Since 1987, the DPR surveillance program has made an effort to collect cases related to antimicrobial/disinfectant products. Before 1987, there had never been more than 50 antimicrobial cases reported in a year. From 1988 through 1991, antimicrobial cases varied from approximately 800 to approximately 900 annually and then began a steady decline. In years without CPCS participation, CBLS supplied 80–98% of antimicrobial cases. In 2007, 34% of antimicrobial reports came from CPCS and 55% from CBLS.

(d) Target Population This program attempts to capture any pesticide-related health problem evaluated by a California physician. Because of reliance on workers’ compensation, occupational exposures are more fully reported than non­ occupational exposures. CACs also investigate complaints registered by citizens and not by physicians. At their discretion, CACs forward to PISP the results of investigations into citizen complaints, and PISP scientists follow explicit standards to determine which to include in the illness surveillance database. (e) Period of Time of Data Collection Annual data are made available approximately 14 months after the end of each calendar year. (f) Periodicity of Reports Reports are released annually. The text of the report and tabulations summarizing the year’s data are posted on the Internet at http://www.cdpr.ca.gov/docs/whs/YYYYpisp. htm (where “YYYY” represents a four-digit year). Reports from 1996 on remain accessible using this naming convention or via the DPR website at http://www.cdpr. ca.gov/docs/whs/pisp.htm. Hardcopy reports are available by request. An online interface at http://apps.cdpr. ca.gov/calpiq was announced with the release of the 2007 annual report. This interface allows any Internet user to query PISP data from 1992 through the most recent year released. (g) Findings PISP records and analyzes data in terms of both “cases” and “episodes.” A case is the PISP’s representation of a pesticide exposure and its apparent effects on one individual’s health. An episode is an incident in which one or more people experience pesticide exposure from a particular source with subsequent development or exacerbation of symptoms. From 1992 through 2007, PISP investigated 18,053 episodes involving suspected pesticide exposures to 26,324 individuals. After evaluating investigation reports, PISP scientists concluded that health effects on 18,125 people were at least possibly attributable to pesticide exposure encountered in 11,946 episodes. These totals include 8 pesticide-associated cases involved in 4 episodes identified in 1991; the 8 cases are excluded from further discussion in this chapter. The following findings apply to the 18,117 cases evaluated as definitely, probably, or possibly attributable to exposure in 11,942 episodes that occurred from 1992 through 2007. Table 61.13 presents demographic characteristics of the affected people and illustrates the preponderance of occupational exposures among case reports. The total annual number of acute pesticide-associated illnesses ranged from 438 to 1856 during this period, with

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Table 61.13  Demographic Distribution of Case Reports Received by the California Pesticide Illness Surveillance Program, 1992–2007, in which Health Effects Were Evaluated, After Investigation, as Definitely, Probably, or Possibly Related to Pesticide Exposurea Occupationalb

Age range

Nonoccupational

Totalc

Male

Female

Unknown sex

Total occupational

Male

Female

Unknown sex

Total nonoccupational

0–4

0

0

0

0

147

125

1

273

273

5–14

7

3

0

10

282

290

11

583

593

15–17

100

68

0

168

52

60

1

113

281

18–24

1713

833

0

2546

133

119

0

252

2798

25–34

2648

1588

2

4238

178

191

0

369

4608

35–44

2021

1646

0

3667

229

241

3

473

4141

45–54

1114

989

1

2104

178

169

0

347

2452

55–64

490

313

0

803

98

118

0

216

1019

65

74

43

0

117

92

105

0

197

315

Unknown

484

443

17

944

281

387

24

692

1637

Total

8651

5926

20

14,597

1670

1805

40

3515

18,117

a

A definite relationship indicates that both physical and medical evidence document exposure and consequent health effects. A probable relationship indicates that limited or circumstantial evidence supports a relationship to pesticide exposure. A possible relationship indicates that health effects correspond generally to the reported exposure, but evidence is not available to support a relationship. b All exposures that occurred while the affected person was at work are considered occupational. Occupational exposures are more fully reported than nonoccupational exposures. c Totals include five cases that could not be characterized as occupational or nonoccupational.

an obvious downward trend over time. Table 61.14 shows the numbers of cases and episodes attributed annually to major pesticide classes. Antimicrobials and insecticides are implicated more often than other types of pesticides. Fumigants are implicated in the fewest episodes but have the largest average episode size. Except for the number of people affected by fumigant exposure, all of the columns in Table 61.14, including number of fumigant episodes, show proportionately similar rates of decline across time. The numbers of people affected by fumigants show increasing volatility, with several recent episodes affecting more than 100 people each. California regulations identify people who work with pesticides as pesticide “handlers.” This designation applies to anyone who mixes, loads, or applies pesticide and also to those who flag for aerial applications and to people who handle potentially contaminated equipment used for pesticide applications. The data summarized here include only eight flagger cases. Applicators are affected in more than half of all handler cases. Case reports on handlers have declined in proportion to the overall decrease in case reports. There is no clear trend in sources of handler exposure. Among agricultural handlers, approximately 40% did not know how they were exposed, and approximately 25% were directly exposed (e.g., by spills or splashes).

Approximately half of nonagricultural handler exposures were direct (e.g., spills or splashes), and almost 25% involved airborne exposure, often to irritant gas produced by antimicrobial pesticides. In addition to the direct exposures mentioned previously, more than 10% of handler exposures (both agricultural and nonagricultural) occurred by direct spray from the application equipment. Table 61.15 presents the impact of pesticide illness in case totals, lost work time, and hospitalization. It shows that herbicides and antimicrobials affect their handlers more often than they affect other people. Fumigants have the greatest tendency to affect nonhandlers, but handlers exposed to fumigants are more likely than others to experience disability or hospitalization. Figure 61.7 provides the distribution of cases by industry category. The agricultural industry accounted for the highest proportion of occupational cases (39%). Table 61.16 lists the pesticides most frequently identified in investigations of illness episodes and the number of people exposed to each. The entry “unknown” identifies cases in which exposure to some pesticide was documented but investigators could not identify the implicated product. This is relatively common in nonagricultural exposures, in which people often have disposed of the pesticide by the time the case is investigated.

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Table 61.14  Summary by Year and Pesticide Category of Reports Received by the California Pesticide Illness Surveillance Program, 1992–2007, in Which Health Effects were Evaluated, After Investigation, as Definitely, Probably, or Possibly Related to Pesticide Exposurea Pesticide category Insecticides

Herbicides

Fungicides

Fumigants

Sanitizers

Totalb

Other/unknown/ combinations

Episodesd

Casese

Episodes

Cases

Episodes

Cases

Episodes

Cases

Episodes

Cases

Episodes

Cases

Episodes

Cases

1992

356

610

94

96

121

169

40

84

602

708

129

183

1340

1850

1993

289

441

69

76

86

92

44

92

514

576

105

159

1104

1436

1994

271

520

64

68

67

106

31

33

467

533

60

69

959

1329

1995

303

570

75

89

77

110

40

105

506

563

96

156

1097

1593

1996

240

571

75

76

72

83

29

71

515

579

104

200

1034

1580

1997

247

445

85

102

75

128

37

84

393

432

83

158

920

1349

1998

161

265

57

59

65

70

22

23

371

468

67

112

742

997

1999

173

327

67

75

55

66

29

222

285

390

80

114

689

1194

2000

177

414

42

47

59

84

20

38

216

246

53

66

566

895

2001

170

212

38

39

30

34

28

46

235

245

34

43

535

619

2002

194

267

60

63

31

40

23

425

293

372

54

149

655

1316

2003

91

109

36

39

40

46

27

258

261

289

47

60

501

801

2004

94

301

40

66

30

35

19

36

287

311

42

85

512

834

2005

60

136

31

34

29

33

19

445

178

187

31

71

348

906

2006

58

91

16

16

24

57

13

74

119

149

32

50

262

437

2007

208

302

36

37

26

30

24

106

309

378

75

128

678

981

3092

5581

885

982

887

1183

445

2142

5551

6426

1092

1803

11,942

18,117

Total a

A definite relationship indicates that both physical and medical evidence document exposure and consequent health effects. A probable relationship indicates that limited or circumstantial evidence supports a relationship to pesticide exposure. A possible relationship indicates that health effects correspond generally to the reported exposure, but evidence is not available to support a relationship. b Ten episodes each included one or more people exposed to pesticides in addition to those encountered by all the people affected. These 10 episodes are counted both in the “other” column and in the column that represents the common exposure. They are counted as single episodes in the “total” column. c All group episode cases are attributed to the year in which each episode was first identified. d An episode is an event in which a single source appears to have exposed one or more people (cases) to pesticides. e A case is the Pesticide Illness Surveillance Program representation of a person whose health problems may relate to pesticide exposure.

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Yearc

Chapter | 61  Surveillance of Pesticide-Related Illness and Injury in Humans

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Table 61.15  Impact of Health Effects Evaluated by the California Pesticide Illness Surveillance Program as Definitely, Probably, or Possibly Related to Pesticide Exposure, 1992–2007a Total cases

No. incapacitatedb

Total days incapacitatedc

No. hospitalizedd

Days of hospitalizatione

1409

339

1360

13

27

Handlers

450

147

593

10

30

Others

616

95

220

5

20

Fieldworkers

100

33

76

2

9

Handlers

388

78

311

1

3

Others

58

5

12

0

0

Fieldworkers

630

140

637

2

5

Handlers

229

45

192

3

20

Others

176

22

89

3

22

Fieldworkers

202

13

42

2

2

Handlers

145

58

271

5

19

Others

1506

26

83

3

17

Fieldworkers

27

5

26

0

0

Handlers

200

40

114

2

5

Others

276

51

147

2

7

Fieldworkers

763

200

579

15

24

Handlers

210

51

269

1

2

Others

357

37

62

2

2

7742

1385

5083

71

214

Handlers

573

116

305

19

85

Others

2533

600

2534

127

536

Handlers

256

42

154

3

6

Others

180

30

82

7

16

Agriculturalf Insecticides Fieldworkers g

Herbicides

Fungicides

Fumigants

Antimicrobials

Other/unknown/ combinations

Total agricultural Nonagricultural Insecticides

Herbicides

(Continued)

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1338

Table 61.15  (Continued) Total cases

No. incapacitatedb

Total days incapacitatedc

No. hospitalizedd

Days of hospitalizatione

Handlers

35

6

12

0

0

Others

113

30

68

1

16

Handlers

53

16

119

1

1

Others

236

34

111

12

66

Handlers

3761

664

2696

19

55

Others

2161

355

1570

43

143

Handlers

81

12

31

3

8

Others

391

52

125

38

180

Total nonagricultural

10,373

1957

7807

273

1112

h

18,117

3342

12,890

344

1326

Fungicides

Fumigants

Antimicrobials

Other/unknown/ combinations

Overall total a

A definite relationship indicates that both physical and medical evidence document exposure and consequent health effects. A probable relationship indicates that limited or circumstantial evidence supports a relationship to pesticide exposure. A possible relationship indicates that health effects correspond generally to the reported exposure, but evidence is not available to support a relationship. b Includes only cases of people who lost at least one full day of work. Excludes cases for which information on disability was not received. c Counts full days lost from work. Disability of uncertain duration is counted as 1 day. d Includes only cases with at least 24 h of hospitalization. Excludes cases for which information on hospitalization was not received. e Counts 24-h periods of hospitalization. Hospitalization of uncertain duration is counted as 1 day. f Agricultural cases are those in which the implicated pesticides were intended to contribute to the production of agricultural commodities (including livestock). g Pesticide handlers are people who mix, load, or apply pesticides; flag for aerial applications; or handle equipment contaminated with pesticides. The totals reported here include only eight flaggers. h Includes two cases that could not be characterized as agricultural or nonagricultural. Neither case involved hospitalization or disability.

Of the 18,117 exposures reported, 5503 involved more than one pesticide. Included among the 5503 are 104 cases of exposure to two products with the same active ingredient. Consequently, cases are counted repeatedly in the Table 61.16 list of pesticides among those exposed to two or more active ingredients, any or all of which may have contributed to symptom development or exacerbation. In particular, most exposures to pyrethrins and pyrethroids also involve exposure to piperonyl butoxide or another synergist. (h) Discussion The California PISP maintains the oldest and largest database of verified information on adverse health effects of pesticides in the United States. It collects data on all types of pesticide products, including antimicrobials. The PISP database can be searched based on dozens of variables, including pesticide identity, type of formulation, toxicity

category, type of health effect, circumstances of exposure, and age and sex of the people affected. The program attempts to capture only those events that result in medical consultation, which provides both a threshold of severity and a preliminary screening by clinical judgment. In circumstances that suggest the existence of a hazard that would justify regulatory response, PISP scientists follow explicit standards that allow database entries to record exposures of individuals who did not seek medical attention. Physician reporting is known to be incomplete because the program receives a high percentage of cases by alternative routes. Augmented by case finding via workers’ compensation, the surveillance program has been instrumental in identifying opportunities for regulatory mitigation of occupational exposures. Poison control facilitation provides a much-needed source of information on domestic exposures. With the assistance of poison control, comparable insights into nonoccupational exposures are anticipated.

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Figure 61.7  Distribution of occupational cases by industry, 1992–2007 (n  14,354).

61.2.4  Bureau of Labor Statistics (a) Description Since the early 1970s, the Bureau of Labor Statistics (BLS) has published annual reports on the number of illnesses and injuries in private industry establishments. Beginning in 1992, the information provided in these reports was enhanced to provide additional information. Among the enhancements was the commencement of reporting of occupational pesticide-related illnesses and injuries. Note that data are available only when the pesticide-related illness or injury resulted in the worker being away from work for one or more days. Estimates of the number of workers with pesticide-related illness or injury that resulted in no lost time are not available. The illness and injury estimates provided in these reports are obtained through an annual survey of employers (BLS, 1998). The survey collects data that employers are required to maintain under the Occupational Safety and Health Act of 1970. The disease estimates provided by the survey are based on a scientifically selected probability sample rather than a census of the entire population. The sample is selected to represent all private industry in the United States. Because the data must meet the needs of participating state agencies, an independent sample is selected for each state. Employers are stratified by their Standard Industrial Classification code and by employment size. Employers are then sampled from these strata. For the strata that contain the employers with the largest employment sizes, the allocation procedure places all of the establishments of the frame in the sample; as employment size decreases, increasingly smaller proportions of establishments are included in the sample. The response rate is generally more than 90%

for sampled establishments. By a weighting procedure, sample units are made to represent all units within a sampling strata. (b) Case Definition Occupational pesticide-related illness and injury cases resulting in days away from work are recorded by employers as required under the Occupational Safety and Health Act of 1970. (c) Data Source An annual survey of employers (see Description). (d) Target Population This survey provides an estimate of the number of serious, nonfatal pesticide-related illnesses and injuries in private industry that involved days away from work. Excluded from the survey are self-employed individuals; farms with fewer than 11 employees (this accounts for approximately 50% of farms; National Agricultural Statistics Service, 2009); employers regulated by other federal safety and health laws (i.e., railroad transportation and coal, metal, and nonmetal mining); and federal, state, and local government agencies. (e) Period of Time of Data Collection Approximately 16 months are required to collect, compile, and publish findings following a given calendar year. (f) Periodicity of Reports Printed reports are published annually. Data are also available on the Internet at http://www.bls.gov/iif.

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Table 61.16  Pesticide Active Ingredients Most Frequently Identified in Case Reports Evaluated by the California Pesticide Illness Surveillance Program as Definitely, Probably, or Possibly Related to Pesticide Exposure, 1992–2007a Agricultural, excluding antimicrobialsb

Nonagricultural, excluding antimicrobials

Antimicrobials

Pesticide

Case count

Pesticide

Case count

Pesticide

Case count

Metam-sodium

842

Chlorpyrifos

392

Sodium hypochlorite

2671

Chloropicrin

571

Unknown

344

Quaternary ammonia

971

Sulfur

474

Diazinon

259

Chlorine

367

Chlorpyrifos

210

Malathion

250

Glutaraldehyde

336

Glyphosate

171

Glyphosate

171

Cyanuric acid

234

Propargite

169

Cyfluthrin

121

Unknown

225

Methamidophos

151

Cypermethrin

101

Calcium hypochlorite

135

Dimethoate

86

Sulfuryl fluoride

95

Phenolic disinfectants

132

Cyfluthrin

82

Propetamphos

78

Sodium chlorite

80

Methyl bromide

69

Permethrin

64

Pine oil

68

Paraquat

57

Copper naphthenate

59

Hydrogen chloride

57

Methomyl

53

Propoxur

57

Peroxyacetic acid

37

Unknown

51

Aluminum phosphide

43

Formaldehyde

29

Sulfur dioxide

49

Resmethrin

42

Trichloromelamine

25

Carbofuran

46

Boric acid

37

Ozone

18

Chlorothalonil

45

Methyl bromide

34

Kathon

16

Ddvp

45

Lambda-cyhalothrin

31

Chlorine dioxide

15

Aluminum phosphide

44

Bifenthrin

28

Halogenated hydantoins

14

Phosphine

37

Metam-sodium

26

Streptomycin

14

Diazinon

36

Acephate

25

Hydrogen peroxide

8

Mevinphos

36

Sulfur

574

Piperonyl butoxide

702

Sodium hypochlorite

381

Chlorpyrifos

565

Pyrethrins

679

Quaternary ammonia

306

Methomyl

318

Synergist

449

Hydrogen chloride

227

Dimethoate

281

Chlorpyrifos

364

Hydrogen peroxide

109

Fenpropathrin

247

Permethrin

196

Peroxyacetic acid

106

Profenofos

244

Petroleum distillates

192

Phosphoric acid

102

Oxydemeton-methyl

223

Allethrin

171

Cyanuric acid

79

Petroleum oil

211

Diazinon

152

Ethyl alcohol

64

Chloropicrin

209

Rotenone

125

Pine oil

57

Myclobutanil

189

Methoprene

120

Phenolic disinfectants

57

Copper hydroxide

181

Cyfluthrin

105

Isopropyl alcohol

56

Imidacloprid

179

Propetamphos

79

Citric acid

55

Methyl bromide

162

Glyphosate

72

Edta

53

Esfenvalerate

156

Tetramethrin

69

Unknown

43

Chapter | 61  Surveillance of Pesticide-Related Illness and Injury in Humans

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Table 61.16  (Continued) Agricultural, excluding antimicrobialsb

Nonagricultural, excluding antimicrobials

Antimicrobials

Pesticide

Case count

Pesticide

Case count

Pesticide

Case count

Chlorothalonil

151

Acephate

59

Sodium carbonate

34

Abamectin

137

Fenoxycarb

57

Sodium metasilicate

27

Glyphosate

130

Cypermethrin

55

Trisodium phosphate

26

Propargite

126

Ddvp

55

Calcium hypochlorite

24

Permethrin

121

Petroleum oil

54

Formaldehyde

23

Diazinon

121

Propoxur

42

Sodium chlorite

18

Iodine complex

18

a A definite relationship indicates that both physical and medical evidence document exposure and consequent health effects. A probable relationship indicates that limited or circumstantial evidence supports a relationship to pesticide exposure. A possible relationship indicates that health effects correspond generally to the reported exposure, but evidence is not available to support a relationship. b Agricultural cases are those in which the implicated pesticides were intended to contribute to the production of agricultural commodities (including livestock).

Figure 61.8  Number of occupational pesticide-related illness cases resulting in one or more days away from work by pesticide functional class, United States, 1992–2007 (data from Bureau of Labor Statistics, 2007).

(g) Findings Between 1992 and 2007, the annual number of pesticiderelated illness and injury cases ranged from 120 to 950 (Figure 61.8). In most of these years, insecticides were responsible for most cases. The exceptions were in 1995 and 1999, when fumigants accounted for the largest number of cases, and 2003 and 2005, when herbicides accounted for the largest number of cases.

(h) Discussion BLS provides data on occupational pesticide-related illness only. In addition, pesticide-related illness data are available only for cases that result in lost work time, suggesting that only the more severe cases are recorded. Data on pesticide-related illness are available beginning in 1992. Because the number of identified cases is relatively small, and because this is a weighted sample and not a census of

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Table 61.17  Counts of Unintentional Pesticide-Related Deaths by Pesticide Class Using Multiple Cause-of-Death Data, 1997–2005a Pesticide-related cause of death (ICD-9 “E” code/ICD-10 code)

1997

1998

1999

2000

2001

2002

2003

2004

2005

Total

Insecticides of organochlorine compounds (E863.0/T60.1)

0

1

1

0

0

2

1

1

0

6

Insecticides of organophosphorus or carbamate compounds (E863.1, E863.2/ T60.0)

2

5

6

1

2

2

5

3

4

30

Other and unspecified insecticides (E863.4/T60.2)

2

2

2

2

7

0

6

6

2

29

Herbicides or fungicides (E863.5, E863.6/ T60.3)

6

2

3

3

3

2

4

4

6

33

Rodenticides (E863.7/T60.4)

4

0

4

2

1

5

2

0

2

20

Fumigants (E863.8)

4

0

—

—

—

—

—

—

—

4

Other and unspecified (E861.4, E863.3, E863.9, E980.7/T60.8, T60.9, X48, Y18)

16

14

12

3

6

5

3

2

4

65

34

24

28

11

19

16

21

16

18

187

Total a

The underlying and all mentioned causes of death were coded using the International Classification of Diseases, 9th revision (ICD-9; World Health Organization, 1977) for years 1997 and 1998 and International Classification of Diseases, 10th revision (ICD-10; World Health Organization, 1992) for years 1999–2005. A separate code for fumigants is not available in ICD-10, and fumigant-related deaths after 1998 are likely placed in the “other and unspecified” category.

the entire population, the estimates have the potential to vary widely from year to year. These limitations may explain the high number of cases in 1995 associated with fumigant exposure. Furthermore, because of the sparse number of identified cases, details on pesticide chemical class and active ingredients, industry and occupation, and circumstances surrounding the pesticide exposure are unavailable.

61.2.5  Vital Status Statistics: Multiple Causes of Death (a) Description The National Center for Health Statistics (NCHS) of the CDC releases a public-use vital statistics tape file for each data year. This file contains a data record for all deaths occurring annually in the United States. Each data record contains the underlying cause of death, other mentioned causes of death, and demographic data. The public-use data files can be purchased from the National Technical Information Service or the Government Printing Office, or they can be downloaded at the Mortality Statistics Branch, Division of Vital Statistics, NCHS website (http://www. cdc.gov/nchs/data_access/VitalStatsOnline.htm). (b) Case Definition Causes of death for the years 1979–1998 were coded using the International Classification of Diseases, 9th revision (ICD-9)

[World Health Organization (WHO), 1977], and for years 1999–2005 the International Classification of Diseases, 10th revision (ICD-10) was used (WHO, 1992). Any of the following ICD-9 causes of death mentioned on the death certificate were included: 989.3, 989.4, E861.4, E863.0, E863.1, E863.2, E863.3, E863.4, E863.5, E863.6, E863.7, E863.8, and E863.9. Any of the following ICD-10 causes of death mentioned on the death certificate were also included: T60.1, T60.2, T60.3, T60.4, T60.8, T60.9, X48, and Y18. The pesticide class that corresponds to each of these codes is provided in Table 61.17. Separate codes are used for suicidal poisonings and poisonings possibly related to suicide. Suicides and possible suicides were excluded from the analyses provided here, and the corresponding codes are not provided. In addition, data on specific pesticide active ingredients are not available. (c) Data Source Multiple-cause-of-death public-use tape files have been released for each data year beginning in 1968. Data from 1997 to 2005 are provided here. (d) Target Population The 50 states, New York City, and the District of Columbia. (e) Period of Time of Data Collection Approximately 28 months are required to collect, compile, and make available data on the website following the end of a given calendar year.

Chapter | 61  Surveillance of Pesticide-Related Illness and Injury in Humans

(f) Periodicity of Reports Public-use data tapes are available annually. (g) Findings Between 1997 and 2005, there were a total of 187 unintentional deaths related to pesticide exposures in the United States. During this 9-year period, the annual number of death certificates that mentioned pesticiderelated illness and injury ranged from 11 to 34 (see Table 61.17). Insecticides were mentioned on 65 death certificates, making this the most common pesticide product type associated with pesticide-related deaths (see Table 61.17). The largest number of insecticide deaths involved organophosphate and carbamate compounds (n  30). The median age of death was 62 years. A total of seven (4%) deaths were among children younger than 10 years. Most of the pesticide fatalities were among those of white race (82%) and male (71%). However, blacks accounted for a slightly disproportionate share of cases (15%). There are only limited data available on the circumstances of these pesticide-related deaths. A total of 66 (35%) occurred in the home, 2 (1%) occurred in a public building, 2 (1%) occurred in industry, 1 (1%) occurred on a farm, 3 (2%) occurred in a residential institution, 2 (1%) occurred on a street or highway, 1 occurred in a recreational setting, and 9 (5%) were noted to have occurred in an “other” location. Data on the location of the poisoning were not available for 101 (54%) of the deaths. (h) Discussion The multiple cause-of-death data are a useful source of data on pesticide-related deaths. However, only the most severe poisonings are included in this data source. There has been little change in the number of pesticiderelated deaths during the past 9 years. Furthermore, they are similar to the numbers reported for the years 1987–1996 (Calvert et al., 2001). However, the numbers are lower than those reported in the 1970s and earlier (Hayes and Vaughn, 1977). Pesticide-related deaths numbered 97 in 1961 and 33 in 1974 (Hayes and Vaughn, 1977). These data also demonstrate declines in the number of pesticide-related deaths among children younger than 10 years. Between 1997 and 2005, these children accounted for 4% of pesticide-related deaths. This is low when considering that children younger than 10 years comprise 13% of the U.S. population (U.S. Census Bureau, 2006). Furthermore, children younger than 10 years accounted for a lower proportion of all pesticide-related deaths during this 9-year period compared to the more distant past. Between 1987 and 1996, children younger than 10 years accounted for 10% of all pesticide-related deaths (Calvert et al., 2001). In 1974, the proportion was 32%,

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and in many of the years before 1973 it was more than 50% (Hayes and Vaughn, 1977). Nonetheless, any childhood poisoning fatality is a tragedy and highlights the need for ongoing efforts to prevent pesticide access among children. It appears that blacks no longer account for a highly disproportionate share of cases. In 2005, 13% of the U.S. population was black (U.S. Census Bureau, 2006), whereas between 1997 and 2005 blacks accounted for 15% of pesticide-related deaths. In contrast, between 1987 and 1996, blacks accounted for 26% of pesticiderelated deaths. The limitations of this data source are many. First, only the most severe (i.e., fatal) cases are included. Second, it is likely that not all pesticide-related deaths were included because some may have been coded to other nonspecific causes of death. For example, Hayes and Vaughn (1977) found that in 1973 and 1974, only 63% of accidental pesticiderelated deaths were coded with the correct “E” code. Finally, details on the circumstances of exposure are not available on the multiple-cause-of-death data file. Collection of such details would require direct queries to the health care provider, as has been done in the past (Hayes, 1976; Hayes and Vaughn, 1977). However, information on each decedent in the multiple-cause-of-death data file includes age, race (including Hispanic origin), gender, state of residence (for pre-2005 data only), injury at work, marital status, and place of the accident.

61.2.6  National Hospital Discharge Studies: Colorado State University (a) Description Three previous studies were conducted by Colorado State University to estimate the nationwide incidence rates for hospitalized acute pesticide poisoning. These studies covered the intervals 1971–1973, 1974–1976, and 1977–1982. No similar national studies have been conducted using data after 1982. We provide information on the most recent study (Keefe et al., 1990). The nationwide poisoning estimates provided by this study are based on a stratified random sampling procedure involving all general care hospitals. States were placed in three strata based on the state’s average rate of hospitalized pesticide poisonings for the years 1971–1976. Hospitals were sampled from each state; however, a higher proportion of hospitals were sampled from those states in the stratum with the highest hospitalized pesticide poisoning rates. Approximately 6% (368 hospitals) of all U.S. general care hospitals were included in the study. (b) Case Definition Hospitalized pesticide poisoning case.

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(c) Data Source A survey of general care hospitals (see Description). For the sampled hospitals, all medical records were reviewed. For patients with designated diagnoses from the International Classification of Diseases, 8th revision, Adapted for Use in the United States (ICDA), medical records were reviewed and appropriate data were abstracted. (d) Target Population This survey provides an estimate of the nationwide incidence of hospitalized pesticide poisoning cases in the United States. (e) Period of Time of Data Collection The last survey covered the time period 1977–1982. (f) Periodicity of Reports Printed reports are available for each of the later two studies (Keefe et al., 1990; Savage et al., 1980), and a review of the findings from the first two studies is also available (Keefe et al., 1985). (g) Findings Between 1977 and 1982 in the United States, the annual number of hospitalized unintentional pesticide poisoning cases was estimated to average 2380 (range, 2127–2991). For this time period, the average annual number of hospitalized intentional pesticide poisoning cases was 454. The estimated annual number of occupational cases averaged 814 (range, 513–1077). The occupations with the greatest number of pesticide poisonings were farmers and commercial applicators. Organophosphates were most often involved in occupational pesticide poisoning cases, accounting for approximately 43% of occupational cases. Following are the top 10 pesticides, listed in order (highest to lowest), responsible for the most number of occupational pesticide poisonings that required hospitalization: parathion, malathion, methomyl, carbofuran, 2,4-D, mevinphos, methyl parathion, disulfoton, aldicarb, and glyphosate (Blondell, 1997). Children 0–4 years old accounted for 57% of all unintentional nonoccupational hospitalized pesticide poisoning cases. (h) Discussion These hospital discharge data were a useful source of data on severe pesticide poisoning cases but are no longer timely. Unfortunately, the most recently available data are from 1977–1982. Funding for this study was cut by the U.S. EPA in the early 1980s due to agency budget cuts and redirected priorities. Furthermore, because the data were obtained by a weighted sample and not by a census of all hospitals, some imprecision likely exists in the estimates.

Hayes’ Handbook of Pesticide Toxicology

61.2.7  National Hospital Discharge Survey: National Center for Health Statistics (a) Description The NCHS conducts an annual survey of nonfederal shortstay hospitals in the United States. The survey has been conducted annually since 1965. Data are collected from a sample of inpatient records acquired from a national sample of hospitals. The data represent a sample of discharges and not patients. Therefore, people with multiple discharges can be sampled more than once. Only general hospitals, children’s general hospitals, or hospitals with an average length of stay of less than 30 days are included. Federal, military, and Department of Veterans Administration hospitals are excluded, as are hospital units of institutions (e.g., prison hospitals) and hospitals with fewer than six beds. The National Hospital Discharge Survey uses a threestage probability sampling design. Approximately 500 hospitals are sampled to acquire approximately 270,000 hospital discharge records. Collected samples are weighted to provide national estimates. Variables available in this data set include age, sex, race, marital status, month of admission, status at discharge, number of days of care, geographic region of the hospital, discharge diagnoses (up to 7) coded according to the ICD-9 categories, and source of payment. Occupation of the patient is not available (NHS, 2009). (b) Case Definition Hospitalized accidental pesticide poisoning case. Any of the following diagnoses are eligible for inclusion (ICD-9): E863.0, E863.1, E863.2, E863.3, E863.4, E863.5, E863.6, E863.7, E863.8, ad E863.9 (the pesticide class that corresponds to each of these codes is provided in Table 61.17). E863 is “accidental poisoning by agricultural and horticultural chemical and pharmaceutical preparations other than plant foods and fertilizers.” Other codes are available to identify poisonings by cleaning products, petroleum products, corrosives, and unspecified solid and liquid substances, but these were not analyzed for this report. Suicides and possible suicides were excluded from the following analysis. (c) Data Source A survey of general-care hospitals (see Description). (d) Target Population This survey attempts to provide an estimate of the nationwide incidence of hospitalized pesticide poisoning cases in the United States. (e) Period of Time of Data Collection Annual surveys have been conducted since 1965. Data were analyzed for 2006, the most recent collected survey.

Chapter | 61  Surveillance of Pesticide-Related Illness and Injury in Humans

(f) Periodicity of Reports Approximately 18 months are required to collect, compile, and publish findings for a given calendar year. Public-use data tapes are published annually. The E-code data are considered too unreliable to be included in the printed annual report. (g) Findings In 2006, there were two discharges with an E863 discharge diagnosis (accidental poisoning by agricultural and horticultural chemical and pharmaceutical preparations other than plant foods and fertilizers). One discharge involved a fungicide (E863.6), and the other involved an unspecified pesticide (E863.4) (NCHS, unpublished data). This sample is too small to provide a weighted estimate of the nationwide incidence of hospitalized pesticide poisoning cases. (h) Discussion The National Hospital Discharge Survey is not a reliable source of data for acute pesticide-related illness and injury. Less than half of the sampled hospitals provide data on E-codes (NCHS, unpublished data). For this reason, E-code data are not published in the annual reports of the National Hospital Discharge Survey. The Nationwide Inpatient Sample (NIS) is another source of hospital discharge data and may contain more complete information on hospitalized pesticide poisoning cases. The 2006 NIS contains all discharge data from 1045 hospitals located in 38 states, approximating a 20% stratified sample of U.S. community hospitals. NIS data are available for purchase through the Healthcare Cost and Utilization Project website (HCUP, 2009).

61.2.8  South Carolina Hospital Discharge Surveys (a) Description The Medical University of South Carolina has periodically conducted a survey of hospitalized pesticide-related illness and injury in South Carolina. The initial survey was published in 1975 and included data from 1971–1973 (Caldwell and Watson, 1975). The latest report contains information on hospitalizations and emergency room visits for pesticide poisoning from 1997 to 2001. All 63 nonfederal hospitals in the state participated in the survey (Simpson et al., 2004). (b) Case Definition Any inpatient medical record that contains one of the following ICD-9 codes: 989.2 (chlorinated pesticides), 989.3 (organophosphate and carbamate pesticides), and 989.4 (other pesticides).

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(c) Data Source A survey of all primary care hospitals in South Carolina that involved a detailed review of case charts by a boardcertified physician. (d) Target Population This survey provides an estimate of the incidence of hospitalized pesticide poisoning cases in South Carolina. (e) Period of Time of Data Collection Periodic surveys have been conducted since the early 1970s. (f) Periodicity of Reports Reports and published papers provide the findings of the periodic surveys. (g) Findings The most recent survey found 148 hospitalized pesticiderelated cases from 1997 to 2001 (mean  29.6 cases/year) (Simpson et al., 2004). Accidental, nonoccupational poisoning in adults and children accounted for 50% of the total number of hospitalized cases, 2% of cases were occupationally related, 40% were intentional ingestions, and 7% could not be classified. There were a total of 3 hospitalized and 9 emergency room visits for occupationally related pesticide poisoning, and 8 of these cases were from the agriculture industry. The ratio of outpatient to inpatient hospital care for pesticide poisonings was 5.8 to 1.0. (h) Discussion The mean number of hospitalized pesticide cases increased since the last published report (Caldwell et al., 1997). The average annual number of hospitalized cases had declined from a peak of 79 cases/year in 1979–1982 to 22 cases/ year in 1992–1996, but it increased to 29.6 cases per year in 1997–2001. The authors attributed the reduced numbers in 1992–1996 to the fact that the participation rate among hospitals was only 90% versus 100% participation in the most recent time interval. An ongoing decline in occupationally related hospitalized cases continued to be observed (from a mean of 20 cases/year in 1979–1982 to 0.6 cases/year in 1997–2001). The authors attribute this decline to the success of pesticide applicator training programs, the licensing and certification of applicators using restricted pesticides, and the increasing use of pesticides with lower toxicity (i.e., pyrethrin/pyrethroid insecticides). Intentional nonoccupational exposures accounted for the highest proportion of pesticide-related hospitalizations (40%). Only the most severe pesticide-related illnesses are captured by hospitalizations. The two most recent surveys also included information on outpatient emergency room visits,

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which generally consist of lower severity cases. From 1997 to 2001, the overall ratio of emergency room visits to hospitalizations for pesticide poisonings was 5.8 to 1, although the rate for occupational pesticide poisonings was only 3:1. This finding could imply that occupational pesticide poisonings are of higher severity or that occupational cases are less likely to seek emergency room treatment.

61.2.9  National Agricultural Workers Survey (a) Description Since the early 1950s, the U.S. government has attempted to monitor the size, composition, and needs of the agricultural labor force. This function was transferred from the U.S. Department of Agriculture to the U.S. Department of Labor (DOL) in 1987, and the National Agricultural Workers Survey (NAWS) reached its present form in 1989. NAWS cooperates with other government agencies by including questions that help measure the needs that can be addressed by migrant education, migrant health, and Census Bureau programs, among others. Since 1993, the U.S. EPA has contributed a limited set of questions designed to elicit information about health effects associated with pesticide exposure. In the October 1998 cycle, the health effects questions were expanded and improved. The number of health effects questions has varied since 1998 and is determined by the amount of survey funding received by DOL. The survey includes questions on medical history, use of medical services, participation in pesticide training, and housing conditions. Although the survey includes questions on general pesticide exposure (e.g., “In the last 12 months, did you load, mix, or apply pesticides?”), information on exposure to specific pesticide products or pesticide classes is not obtained. A complete occupational history for the year preceding interview is also obtained. (b) Data Source Respondents for this annual survey are selected by a multistage stratified process. The program defines 12 regions, each of which is further divided into several farm labor areas. Each farm labor area is an aggregate of counties that have similar agricultural and economic characteristics. At least three farm labor areas are selected from each region. One county is then selected from each selected farm labor area. Staff then compile lists of all farms within the selected counties and solicit cooperation from a random sample of the farms. At participating farms, the employees are sampled with a probability proportional to the square root of the size of the farm workforce. By sampling and recruiting workers at their worksite, this survey minimizes the undercounting of this population. After they are identified and recruited, the farm workers are interviewed face-to-face outside of working hours at

Hayes’ Handbook of Pesticide Toxicology

home or another non-workplace location. The interview lasts approximately 1 h. (c) Target Population This survey provides national estimates on the U.S. crop labor force (note that workers involved in livestock production are excluded). (d) Period of Time of Data Collection The survey interviews approximately 500–1500 agricultural workers in each of three annual interview cycles. The number interviewed in each cycle depends on data needs and funding availability. Cycles begin in February, June, and October and last 15 or 16 weeks. These cycles were selected to reflect the seasonality of crop production and employment. (e) Periodicity of Reports Periodic reports presenting aggregate findings are available from the Office of the Assistant Secretary for Policy of the DOL. Public-use data files are also available at the NAWS website. Finally, if the desired data are not available in published reports or in the public-use files, specialized analyses can be requested from NAWS staff. Additional information on NAWS is available at its website (http:// www.doleta.gov/agworker/naws.cfm). (f) Findings In 1999, with funding provided by NIOSH, NAWS included questions to determine if crop workers were poisoned by pesticides. This information was collected in two parts. First, NAWS asked crop workers if they were exposed to pesticides by “having them sprayed or blown on you,” “spilled on you,” or “when cleaning or repairing containers or equipment used for applying or storing pesticides.” NAWS then asked if the crop workers became “sick or [had] any reaction because of this incident.” Analyses of these data found that 3.2% of crop workers acknowledged exposure during the previous 12 months, of whom 43.4% reported getting sick or having a reaction (Calvert et al., 2008). That is, 1.4% of U.S. crop workers attributed health effects such as skin problems (59%), eye problems (24%), nausea/vomiting (30%), headache (26%), and numbness/ tingling (12%) to pesticide exposure during the preceding 12 months. In a separate NAWS question, 0.6% of all crop workers reported that in the past 12 months they had “received medical attention by a doctor or nurse due to pesticide exposure.” To our knowledge, neither these nor similar questions to assess the incidence of pesticide poisoning were included in NAWS surveys before or after 1999. These estimates are much higher than other estimates. For example, an analysis of acute pesticide poisoning data from SENSOR-Pesticides and CDPR found an average annual acute occupational pesticide poisoning

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incidence rate of 0.05% among all agricultural workers and 0.07% among farm workers. (g) Discussion NAWS can provide a useful source of data on acute pesticide poisoning among U.S. crop workers. Because it is a survey, it circumvents limitations found with surveillance systems that rely on physician reports. Physician reporting requires that the affected individual seeks medical care, that a diagnosis of pesticide-related illness or injury is made or suspected, and that the physician is aware of the need to report the suspected case. However, the NAWS findings on acute pesticide poisoning can be affected by whether the respondents correctly determined that they were poisoned and whether they were willing to report their status. It would be helpful if future NAWS efforts repeated the pesticide poisoning questions from 1999. This would allow an examination of trends in acute pesticide-related illness among U.S. crop workers and would be helpful to assess the effectiveness of intervention efforts to prevent these illnesses.

61.2.10  Surveillance Efforts of International Organizations Pesticide poisonings are a major concern throughout the world. The problem may be even greater in developing countries because of the impracticality of much personal protective equipment in humid tropical areas, because farmers are often illiterate, because the pesticide label is often not available in the local language, and because of employer disregard for worker health and safety (Figure 61.9) (Eddleston et al., 2002). WHO estimates that there are up to 5 million acute unintentional pesticide-related illnesses and injuries per year, and that annually there are 20,000 deaths related to unintentional pesticide poisoning (Levine and Doull, 1992). Although it is well recognized that acute pesticide poisoning is a major public health problem in developing countries, surveillance of this condition in developing countries is scarce. The data available are not adequate to address the nature of the problem and are usually limited to ad hoc studies that are neither compatible nor comparable with each other, making estimates and evaluations difficult to undertake. To overcome this and assist developing countries in sound management of pesticides, efforts have been undertaken by some international organizations.

61.2.10.1  World Health Organization In 1972, a WHO expert committee made the first global estimate of the number of cases of acute pesticide poisoning, on the basis of a theoretical model. Although these numbers raised some controversies, the information

Figure 61.9  A youth in Peru applying pesticides (courtesy of David L. Parker, MD, MPH).

available showed that the problem existed, was important, and needed attention. The International Program on Chemical Safety (IPCS) initiated a range of activities designed to characterize the true extent and severity of pesticide poisoning worldwide and set up the basis for surveillance systems. This included development of harmonized methods for governments to collect data on acute pesticide poisonings. A project titled “Epidemiology of Pesticide Poisoning – Harmonized Collection of Data on Human Pesticide Exposures” was launched. A common set of tools included a pesticide exposure record (Annex I) and pesticide poisoning severity score (PPSS) that was adapted from the poisoning severity score (PSS) developed by IPCS in collaboration with the European Commission and the European Association of Poison Centres and Clinical Toxicologists (Persson et al., 1998). Five countries (India, Indonesia, Myanmar, Nepal, and Philippines) participated in the project by collecting data using the developed tools and training medical staff on the diagnosis and treatment of pesticide poisonings. The project demonstrated the utility and acceptance of the tools for acute pesticide poisonings treated at hospitals, but it did not appear to reflect the situation concerning occupational and accidental exposures. It was recognized that population-based studies are required to collect information about cases that are not captured in hospital records. As a continuation to the project, a surveillance protocol was developed to collect data on pesticide poisonings at the tertiary, secondary, and primary health care levels. The WHO/IPCS Pesticide Advisory Group that met in Washington in 2002 agreed on the protocol for the development of a national plan for the surveys. The acute poisoning protocol included, but was not limited to, an assessment of data availability, a sampling strategy for populationbased surveys, an interviewing protocol for surveys, a standard case definition for acute pesticide poisoning, and standardization of poisoning severity based on a simplified version of the PPSS (WHO, 2002). Unfortunately, due to

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financial constraints, it was not possible for WHO to test the methodology. The Pan American Health Organization, through its Division of Health and Environment, implemented a project (PLAGSALUD in Spanish) that was financed by the Danish Agency for International Development on the occupational and environmental aspects of exposure to pesticides in the Central American subregion (including Belize, Costa Rica, El Salvador, Guatemala, Honduras, Nicaragua, and Panama) (Henao and Arbelaez, 2002). The 10-year project, which was initiated in 1994, aimed to significantly reduce the health problems related to pesticides and support the implementation of sustainable agriculture alternatives. Technical cooperation was provided in epidemiological surveillance, research, education, interinstitutional coordination, and the strengthening of legislation. One of the main results of this project was incorporation of surveillance of acute pesticide poisoning in the surveillance systems of the seven countries and also improvement of reporting in epidemiological surveillance. The incidence rate of acute pesticide poisoning in the Central American subregion increased from 6.3 per 100,000 population in 1992 to 19.5 in 2000, but this was most likely due to an increase in surveillance efforts by the PLAGSALUD project. The mortality rates also showed a rising trend in the same period, with a risk of death of 0.3 per 100,000 population in 1992 increasing to 2.1 per 100,000 in 2000. These could be related to better surveillance and a greater awareness of pesticide poisoning by medical personnel. The study on underreporting conducted in the context of the PLAGSALUD project in 2001 showed that underregistration was still a problem, ranging from 80 to 99% in six countries where community surveys were conducted (Henao and Arbelaez, 2002).

61.2.10.2  Food and Agriculture Organization of the United Nations Pesticide management is carried out within the overall framework of the Plant Production and Protection Division of the Food and Agriculture Organization (FAO). It is designed to work together with member countries and other international organizations as a partner to introduce sustainable and environmentally sound agricultural practices that reduce health and environmental risks associated with the use of pesticides. The International Code of Conduct on the Distribution and Use of Pesticides stipulates governments “to carry out health surveillance programs of those who are occupationally exposed to pesticides, and investigate, as well as document, poisoning cases”(Article 5.1.3) and “provide guidance and instructions to health workers, physicians, and hospital staff on the treatment of suspected pesticide poisonings”(Article 5.1.4) (FAO, 2002). The Regular Monitoring Report Form of the Guidelines on Monitoring and Observance of Code of Conduct asks

Hayes’ Handbook of Pesticide Toxicology

governments to provide information on occupational exposure to pesticides and poisonings (FAO, 2006). Within the context of the FAO Program for Community IPM in Asia, a field document was prepared in 2001 to pilot a farmer-based surveillance system in a community in Vietnam and Thailand. Each farmer was asked to fill out a form each time he or she sprayed. The information recorded included name, gender, address, date and spray event number, list of pesticides used, and home treatment used. Any sign or symptom experienced during or up to 24 h after spraying was circled on a body map that showed 29 potential signs and symptoms associated with pesticide poisoning (Murphy, 2002). For 12 months, 50 farmers in northern Vietnam recorded after every spraying session any adverse health effect and the pesticide that was used. Data were also gathered from 50 controls. Of the 1798 recorded spray operations, 8% were asymptomatic, 61% were associated with vague ill-defined effects, and 31% were accompanied by a least one clear symptom of poisoning. After 6 months of submitting self-reports and receiving feedback, the farmers in the intervention group had a significant reduction in self-reported spraying frequency and use of highly hazardous products (Ia/Ib) compared to the controls, as had their reports of adverse health effects of moderate severity (Murphy et al., 2002).

61.2.10.3  Intergovernmental Forum on Chemical Safety The Intergovernmental Forum on Chemical Safety (IFCS) was created in 1994 and provides an open, transparent, and inclusive forum for discussing issues of common interest and also new and emerging issues in the area of sound management of chemicals. IFCS plays a unique multifaceted role as a flexible, open, and transparent brainstorming and bridge-building forum for governments, intergovernmental organizations, and nongovernmental organizations including those from the private sector. At the third forum of IFCS that was held in Bahia, Brazil, in 2000, one of the goals was “a report will have been prepared on the problem of acutely toxic pesticides and severely hazardous pesticide formulations and recommending sound management options.” At IFCS Forum IV held in Thailand in November 2003, the governments made a series of recommendations on acutely toxic pesticides that included a recommendation to establish or enhance comprehensive national systems for surveillance and reporting of poisoning incidents affecting workers and communities (IFCS, 2003). Many developing countries lack the resources to establish and maintain the necessary surveillance programs and to obtain confirmatory laboratory testing for all possible cases of acute pesticide poisoning. Taking the WHO Pesticide Project Surveillance Working Group case definition (WHO, 2002) as a starting point, a paper was prepared by IFCS on a proposed

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classification tool for acute pesticide poisoning (Thundiyil et al., 2008). The aim was to provide a standard definition classification scheme for acute pesticide poisoning to enable its identification and diagnosis, especially at the field level, in rural clinics, and in primary health care systems. A standardized case definition and classification scheme for acute pesticide poisoning into categories of probable, possible, or unlikely/unknown is proposed. Based on the proposed criteria, laboratory confirmation is not absolutely necessary to meet the standard of a probable acute pesticide poisoning.

61.3  U.S. Environmental protection agency regulations The U.S. EPA is responsible for implementing several regulations that promote the safe use of pesticides and that facilitate surveillance of pesticide-related injury and illness. These regulations are discussed in the context of collecting information about acute adverse effects of pesticides and the regulatory programs available to implement risk mitigation.

61.3.1  The Federal Insecticide, Fungicide, and Rodenticide Act The U.S. EPA regulates the use of pesticides in the United States under the authority of the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA) (7USC§136). FIFRA was originally enacted in 1947 and underwent a major revision in 1972. No pesticide may legally be sold or used in the United States unless it bears a U.S. EPA registration number. It is a violation of the law for any person to use a pesticide in a manner inconsistent with its label. FIFRA gives the U.S. EPA the authority and responsibility for registering pesticides for specified uses, provided that such uses do not pose an unreasonable risk to human health or to the environment. In 1996, FIFRA was amended by the Food Quality Protection Act (FQPA), which included stricter pesticide safety standards, with the intent to better protect infants and children from pesticide hazards. FQPA required that there be a reasonable certainty that residues that result from use of a pesticide in or on any food pose no harm to human health. FIFRA provides a number of remedies that can reduce and mitigate risks from pesticides. If subsequent information indicates that the use of a pesticide would pose unreasonable risks (pertaining to non-food-use pesticides) or a reasonable certainty or risk (pertaining to pesticide residues on food), the U.S. EPA has the authority to suspend or cancel its registration. If a pesticide warrants special handling because of its toxicity, it may be classified for restricted use. Pesticides with a restricted-use classification

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can be applied only by a certified applicator or under a certified applicator’s direct supervision. States administer the certification programs, which require the certified applicator to demonstrate competency with respect to the use and handling of pesticides. The U.S. EPA also delegates to the states the responsibility and authority for enforcement of FIFRA (e.g., investigating and issuing penalties for a label violation). Every U.S. EPA registered pesticide product has a pesticide label. The pesticide label provides directions on how to use a pesticide product and information on where it can be used, application rates (per treatment and per year), frequency of application, need for any required application equipment, and which pests can be controlled. The U.S. EPA classifies all pesticides into one of four acute toxicity categories based on established criteria (40 CFR Part 156). Toxicity is determined by acute animal tests for oral, dermal, inhalation lethality, and corrosive effects to the skin and eyes. Those pesticides with the greatest toxicity are placed in Toxicity Category I. Other pesticides are placed in the remaining three toxicity categories (Categories II–IV). A hazard signal word indicates the toxicity category of a pesticide product. The most hazardous pesticides (i.e., Toxicity Category I) are labeled “Danger,” those with moderate toxicity (i.e., Toxicity Category II) are labeled “Warning,” and less toxic pesticides (i.e., Toxicity Categories III and IV) are labeled “Caution.” Precautionary statements on the label describe the protective clothing and other equipment that must be worn and used when handling or applying a specific pesticide product. They also specify the hazards to humans, children, domestic animals, and the environment. A statement of practical treatment may advise on the signs and symptoms of poisoning, provide information on first aid and antidotes, and provide a note to physicians on appropriate treatment. The label also specifies directions for safe storage and disposal. The label may have a number of statements designed to reduce risk in addition to the requirements listed previously. The label must specify any restrictions on use for factors such as weather, time of day, season of the year, contamination of sensitive areas, and exposure of nontarget species. The label will indicate when an application requires use of enclosed tractor cabs, closed mixing/loading systems for liquids, ventilation, or mechanical flagging devices. Certain more hazardous application methods (e.g., air blast spraying) may be prohibited. Hygiene statements may require washing clothing after the application or prohibitions against wearing contaminated clothing the next day. After application, a restricted entry interval may be imposed whereby unprotected people are not permitted in the treated area for a time period specified on the label. Posting and/or notification of a treated area may be required to warn bystanders and others not directly involved in the application. Preharvest intervals may also

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be required (i.e., the time period before harvest when use of the pesticide product is prohibited). Formulation and packaging requirements can also be imposed under FIFRA to reduce pesticide risks. For example, the container size or percentage active ingredient may be limited. In addition, formulations may require readyto-use solutions instead of concentrates, warning odors or dyes, or a bitter taste to discourage ingestion. Formulations may be limited to types (e.g., dry flowables) that limit exposure to handlers. Packaging design (e.g., water-soluble packets and lock-and-load container design) can also minimize handler exposure. Depending on acute toxicity, the pesticide may be required to be in child-resistant packaging.

61.3.2  Federal Reporting Requirements for Risk Information Section 6(a)(2) of FIFRA requires pesticide registrants to submit to U.S. EPA information concerning adverse effects of their pesticide products. The purpose of the Section 6(a)(2) requirement is to help ensure that U.S. EPA pesticide registration decisions, as well as the terms and conditions of registration, were correct and that a pesticide can be used without posing unreasonable adverse effects to human health and the environment. Information submitted under this rule involves any and all toxicological and ecological studies, antimicrobial product efficacy failure data, and incident reports of injury or illness associated with use of the pesticide product. Incident reports may involve humans, domestic animals, wildlife, plants, surface or groundwater contamination, or property damage. Some registrants have elected to use the services of poison control centers to handle inquiries about adverse effects incidents concerning their products. When the U.S. EPA learns from a third party that a registrant knew about but did not report appropriate risk information, fines of several thousand dollars per case have been imposed on the guilty registrant. For incident reporting of serious or rare incidents, the regulations specify that detailed information must be provided. For example, for more serious human incidents, documentation is requested on the pesticide agent, the circumstances of exposure, and evidence of the type and severity of adverse effects. Exposure circumstances include information on how the exposure occurred, the site where the pesticide was used, the situation, documentation of factors that may have contributed to exposure (e.g., early field reentry), and any other evidence that the label directions were not followed, if available. Adverse effects information includes the route of exposure, list of signs and symptoms, results from medical laboratory tests, type of medical care sought (i.e., none, clinic, hospital emergency department, private physician, poison control center, or hospitalization), time between exposure

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and onset of symptoms, and estimated duration or dose of exposure, if available. Common or minor incidents, on the other hand, can be summarized as counts by product or active ingredient. There are some exceptions to the Section 6(a)(2) reporting requirement. Incidents are not required to be reported if facts establish that the exposure or the effect did not occur, if the registrant has been granted a written waiver from the U.S. EPA, or if the incident involved only minor skin/eye effects from residential use and a warning of such effects is provided on the pesticide label. There are limitations related to the data submitted under the Section 6(a)(2) reporting requirement. Incident reports often consist of anecdotal reports or unproven allegations. Pesticide incidents must be reported even if the pesticide registrant does not think the adverse effect was caused by the pesticide product. Furthermore, pesticide registrants are not required to investigate incident reports. There are no criteria that must be met with respect to certainty of exposure, certainty of reported health effects, or certainty of the toxicological link between the exposure and health effects. However, Section 6(a)(2) data can be useful for identifying emerging pesticide problems, especially if several reports are received relating specific adverse effects to a specific pesticide product. Section 6(a)(2) is not available on the web but can be requested from the U.S. EPA.

61.3.3  National Pesticide Information Center The National Pesticide Information Center (NPIC) provides information about pesticides and pesticide-related topics, including antimicrobial products, using a toll-free telephone service available to any caller in the United States, Puerto Rico, or the Virgin Islands. It also provides a wealth of pesticide-related information on its website. NPIC is funded by the U.S. EPA to provide objective, sciencebased information about a wide variety of pesticide-related subjects, including pesticide products, recognition and management of pesticide poisoning, toxicology, and environmental chemistry. The service can provide chemical, health, and environmental information on more than 1000 pesticide active ingredients incorporated into more than 16,000 different products registered for use in the United States. NPTN operates 7 days a week, excluding holidays, from 6:30 am to 4:30 pm Pacific time. NPTN can be reached by telephone at 1-800-858-7378 or at their website at http://npic.orst.edu.

61.3.4  Worker Protection Standard Recognizing the need for increased worker protection from pesticide exposures, the U.S. EPA promulgated rules in 1974 known as the Worker Protection Standard for Agricultural Pesticides (WPS) (40 CFR 170). The aim of

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WPS is to reduce pesticide exposures among agricultural workers. Workers employed at farms, forests, nurseries, and greenhouses are covered by WPS. By 1992, the U.S. EPA estimated that hired farm workers alone experienced up to 10,000–20,000 illnesses and injuries from pesticide exposures each year (U.S. EPA, 1992) and concluded that the WPS was inadequate in its requirements and scope of coverage. That year, the U.S. EPA revised and expanded the WPS rules to include changes in labeling, coverage of more workers and agricultural operations, prohibition of employer retaliation against workers attempting to comply with the standard, and the following requirements: notification of workers about pesticide applications; restriction of reentry into pesticide-treated areas; and provision of PPE, decontamination supplies, emergency assistance, and pesticide safety training.

61.4  Evaluating surveillance systems The purpose for evaluating surveillance systems is to ensure that available surveillance resources and funding are directed at important public health problems and to determine whether the surveillance systems are operating efficiently and effectively. Guidelines for evaluating surveillance systems are available (CDC, 2001c) and are briefly summarized here. Public health surveillance systems should be evaluated periodically, and the evaluation should include recommendations for improving the systems’ quality, efficiency, and usefulness. When evaluating surveillance systems, it is important to describe the public health importance of the health event of interest. The information provided previously in this chapter supports the public health importance of acute pesticiderelated illness and injury. Surveillance data indicate that a large number of cases occur annually, some of which are fatal. In addition, epidemiologic data suggest that acute poisoning is associated with long-term health effects (Rosenstock et al., 1990; Savage et al., 1988; Steenland et al., 1994). Pesticides are toxic substances and are used in large quantities in both the indoor and the outdoor environment (Kiely et al., 2004). Because society allows pesticides to be disseminated into the environment, society also incurs the obligation to conduct surveillance on the health effects of pesticides. Finally, acute pesticide-related illness and injury are preventable through regulation and enforcement, appropriate training, and by carefully following the instructions on the pesticide label. Data from public health surveillance systems can be used to guide the planning, implementation, and evaluation of these measures. It is also important to assess the usefulness of surveillance systems. This can include describing the actions taken as a result of using data from the surveillance system (e.g., policy changes, regulatory changes, and clinical

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practice changes) and describing other anticipated uses of the data (e.g., detecting disease magnitude and trends, detecting new pesticide hazards or new populations at risk, detecting epidemics, stimulating epidemiological research, and assessing the effectiveness of interventions). The attributes of the surveillance system should also be evaluated. Because some attributes can conflict with other attributes (i.e., excelling in one attribute may hamper the ability to satisfy another attribute), it is important to identify and strengthen those attributes that are most important to a particular surveillance system. It should be recognized that it may not be possible to fully achieve the less important attributes. The attributes that should be evaluated are as follows: Sensitivity: What proportion of the total number of cases or total number of outbreaks is identified by the system? Sensitivity is influenced by whether cases seek medical care, whether those who seek care are correctly diagnosed, and the likelihood that the correctly diagnosed case will be reported to public health authorities. Surveillance conducted through surveys of individuals at risk of pesticide poisoning (e.g., farm workers) can be affected by whether subjects can correctly determine that they were poisoned and whether they are willing to report their status. Substantial resources may be required to evaluate this attribute (e.g., extracurricular efforts to determine the annual incidence of the condition in the community). Systems without high sensitivity can be useful for monitoring trends, as long as the sensitivity remains relatively constant. Flexibility: How adaptable is the system to changing needs or operating conditions? A flexible system can handle changes in case definitions, reporting sources, data elements, and outcomes/diseases/exposures. Flexibility is a characteristic met by many pesticide poisoning surveillance systems. Simplicity: This refers to the structure of the system and its ease of operation. The surveillance system should be as simple as possible. There are several measures to be considered. Is the case definition easy to apply? Are there multiple reporting sources? How much time is spent collecting data? What is the mechanism for transmitting case information/data? How extensive are the staff training requirements? What type of data analysis is required? Who are the users of the data and what is the mechanism for distributing reports/data? Due to the complexity and large number of pesticides, pesticide poisoning surveillance systems are rarely simple. Data quality: This reflects the completeness and validity of the data being collected by the pesticide poisoning surveillance system. It can be assessed by examining the proportion of cases with “unknown” information for various variables or data elements that the surveillance system attempts to capture. It is influenced by

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two other attributes: the sensitivity and positive predictive value. It is also influenced by the level of training of surveillance system staff and the care taken in data management. Data quality impacts two other attributes: acceptability and representativeness. High data quality will improve the acceptability of the surveillance system and should also improve its representativeness. Acceptability: This attribute assesses the willingness of individuals and organizations to participate in the system. It can be assessed by examining the participation of various potential reporting sources, interview completion rates, extent of missing data, and timeliness of reporting. Several factors influence acceptability. What is the public health importance of the health event of interest? Is there timely recognition of a stakeholder’s contribution? Are the time and personnel costs onerous? Have reporting laws been enacted? Does the system provide high-quality, useful information? Positive predictive value: It is important to maximize the proportion of cases reported to the system that actually have the condition. This is because system resources are used to confirm and investigate cases reported to the system. A system with a low predictive value for reported cases suggests that resources may be wasted when those cases are investigated. Inappropriate outbreak investigations may be conducted if a high number of false positives are reported. The predictive value is related to the sensitivity and specificity of the case definition and the prevalence of the condition in the population. Increased specificity and prevalence leads to an increased positive predictive value. Representativeness: This assesses whether findings from the surveillance system can be generalized to the entire target population. It is also important to attempt to identify any population subgroups that may be systematically excluded from the surveillance system. Because surveillance data are often used to calculate morbidity and mortality rates, thought should be given to ensure that the denominator, which is often obtained from a different source (e.g., the Current Population Survey, which consists of data collected by the U.S. Census and the Bureau of Labor Statistics), is comparable with respect to the demographics of the surveillance data in the numerator. As with sensitivity, substantial resources may be required to assess this attribute. Special studies may be useful that seek to identify all cases and then compare them to those cases reported to the system. Timeliness: This refers to the time interval between each step in the surveillance system. Among the more important time intervals to assess are the length of time between an event and its being reported to the surveillance system, the time required to identify trends and outbreaks, and the time required to institute interventions. The importance of timeliness depends on the

urgency of the public health problem and the availability of effective control measures. The pervasive use of computer technology and the Internet, and growing use of electronic health information and electronic medical records, holds promise for improving this attribute. Stability: This attribute refers to the availability and reliability of the public health surveillance system. Measures can include assessing how often the surveillance system is fully operational and the difference between the desired amount of time and actual amount of time required to collect, receive, manage (including data entry and edit checks), and disseminate data. A lack of sufficient resources can affect stability. Unreliable and unavailable surveillance systems can hamper public health interventions. (a) Discussion When evaluating a surveillance system, conclusions and recommendations should be provided. An assessment should be made as to whether the surveillance system should be continued (i.e., Is the health condition under surveillance important? Can justification be made for the resources used by the system?). If it is to be continued, the need for any modifications to the system should be identified. Finally, when making recommendations for modifications, it is prudent to recall that the costs and attributes of the system are interdependent. Improvements in many of the attributes (sensitivity, representativeness, timeliness, data quality, and stability) will likely increase the costs of the surveillance system. In addition, improvements in one attribute may affect performance of another attribute. For example, improvement in positive predictive value may compromise sensitivity and may reduce simplicity. Therefore, these consequences should be considered when recommending modifications.

61.5  Case definition for acute pesticide-related illness and injury A public health surveillance system depends on a clear case definition for the condition under surveillance. It is used to identify individuals with a health outcome of interest. The standardized case definition improves specificity and allows for comparability across geographic areas, time periods, and various reporting sources. It is needed both in epidemiologic studies and to conduct surveillance. The case definition for acute pesticide-related illness and injury can be simple or complex. Those that are used in the surveillance systems described previously are not identical but vary across the systems. In some instances, the clinical diagnosis may be the basis of the case definition. For example, conducting surveillance of unintentional

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pesticide-related fatalities using vital status statistics involves identifying death certificates that contain ICD-9 E codes or ICD-10 codes specific for accidental pesticide poisoning. These ICD-9 and ICD-10 codes are supplied by the health care professional who completed the death certificate. Similarly, data from BLS and AAPCC often involve medical outcome data supplied by the health care provider. Some case definitions may make a determination that differs from the clinical diagnosis. This may occur when the case definition provides guidelines for assessing the certainty of the evidence regarding exposure and health effects. An example of this is the case definition for acute pesticiderelated illness and injury developed for the National Public Health Surveillance System (NPHSS). In contrast, the purpose of the clinical diagnosis is to guide the immediate treatment course for an ill individual. In addition to employing much of the same information used by a surveillance system, a clinical diagnosis often also involves a subjective understanding of a patient’s sensitivity to exposure. This subjective understanding may not be available for consideration by the surveillance system.

61.5.1  The National Public Health Surveillance System Case Definition The NPHSS is a conceptual framework for public health surveillance based on a consensus of practicing epidemiologists at the local, state, and national levels (Meriwether, 1996). Goals of the NPHSS include prioritizing surveillance activities and securing the necessary resources to conduct these activities. Acute pesticide-related illness and injury is one of the conditions identified for inclusion in the NPHSS. The acute pesticide-related illness and injury case definition for the NPHSS was developed using a modified nominal group process (Jones and Hunter, 1995). The group consisted of experts from federal agencies [NIOSH, U.S. EPA, and National Center for Environmental Health (NCEH)], nonfederal agencies (Council of State and Territorial Epidemiologists and Association of Occupational and Environmental Clinics), and state health departments or other state designees. Prior to the first meeting of the group, a proposed case definition was distributed. During the first meeting in September 1995, the case definition was reviewed and revisions were made. Following the meeting, a revised case definition was provided to each of the participants, along with a classification exercise that consisted of three “test” cases that each participant classified using the case definition. The classification exercise identified the need for several modifications to the case definition. Additional meetings held in April 1996 and November 1997, two subsequent classification exercises, and additional iterations via e-mail continued this process until consensus was achieved.

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Because public health agencies seek to prevent all adverse effects from regulated pesticides, the case definition is intended to be applied to any acute adverse health effect resulting from exposure to a pesticide product, including health effects due to an unpleasant odor, injuries from explosion of the product, allergic reactions, and effects associated with inert ingredients. The case definition requires the collection of information in three areas: pesticide exposure, health effects, and evidence supporting a causal relationship between exposure and effect. A case of pesticide-related illness or injury is classified as being definite, probable, or possible. The specific classification category is chosen depending on the level of certainty of exposure, whether health effects were observed by a health care professional, and whether there is sufficient toxicologic information to support a causal relationship between the exposure and health effects. The cases classified into these categories meet the following criteria: Documentation of two or more new adverse health effects that are temporally related to a documented pesticide exposure l Consistent evidence of a causal relationship between the pesticide and the health effects based on the known toxicology of the pesticide from commonly available toxicology texts, government publications, information supplied by the manufacturer, or two or more case series or positive epidemiologic investigations. l

When insufficient toxicologic information is available to determine whether a casual relationship exists between the pesticide exposure and the health effects, a case is classified as “suspicious.” This category is assigned when minimal information is available on the human health effects that can be produced by the pesticide in question (e.g., when there are less than two published case series or positive epidemiologic studies linking health effects to the putative exposure agent). When convincing evidence for an exposure–health effect relationship is not present, the case is classified as “unlikely.” A classification of “not a case” is assigned when there is strong evidence that no pesticide exposure occurred, when no new postexposure signs or symptoms were reported, or when there is definite evidence of a nonpesticide causal agent. The case definition is complex. It requires knowing how to obtain information on each of the following three criteria: exposure (i.e., knowing what environmental and medical tests should be conducted and what questions to ask), health effects (i.e., the ability to review medical records and to solicit a medical history), and the causal relationship (i.e., knowledge about how to find and use appropriate toxicological and medical references). It also requires knowledge and experience with assessing whether the exposure was sufficient to produce the observed health effects. Because of the skills, knowledge, and experience that are required to use this case definition, it is likely that for any

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given case, the extent of agreement among raters on a classification category for a given case will not be total. There are several reasons for the complexity of the case definition. One is that there is no “gold standard” for identifying pesticide-related illness and injury (i.e., there is no symptom, sign, or test that is definitive or pathognomonic for this condition). Therefore, identifying cases of pesticiderelated illness and injury requires assessment of available information on exposure, health effects, and causal relationship. Unfortunately, because there is no gold standard, it is difficult to determine the case definition’s sensitivity, specificity, and positive predictive value. Likewise, it is difficult to assess the degree of misclassification that arises by using the case definition. However, we think most would argue that this case definition reduces misclassification compared to other less rigorous definitions. Another reason for the complexity of the case definition is that it covers all pesticide chemical classes and active ingredients (i.e., literally hundreds of different chemicals). This allows the case definition to be flexible. This case definition is also resource intensive. It requires having trained staff to collect and assess the information needed for case classification. It also requires staff to code and key the data into a database so that the data on each criteria can be organized and analyzed. A major strength of the standardized case definition and the standardized variables (described previously) is that they allow data from participating surveillance systems to be compared and aggregated. This aggregation has permitted the enhancement of knowledge about acute pesticiderelated illness and injury. With this knowledge, the goal of surveillance can be and has been realized: targeting public health resources toward the prevention of acute pesticiderelated illness and injury. In conclusion, a case definition is needed to identify individuals with pesticide-related illness and injury. Currently, the case definition varies across surveillance systems. Where adequate resources exist, it is recommended that the case definition for the NPHSS be used. Regardless of the case definition that is used, all serve the purpose of identifying cases so that appropriate interventions can be targeted.

61.6  Limitations of pesticide poisoning surveillance data When examining surveillance data, one needs to be mindful of its limitations. A discussion of these limitations follows.

61.6.1  Denominators A denominator is needed to calculate rates. Comparing the rate of the condition across different groups is needed to identify high-risk populations and to evaluate risk factors.

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Counts alone of a condition’s occurrence may have little value for identifying disease risk factors. The difficulty of finding appropriate denominator information is one of the most obvious limitations of surveillance data. National populations provide one type of denominator for surveillance data. The absolute counts reveal striking differences in pesticide mortality between industrialized and developing nations. Adjusting for relative population size, through the use of rates, emphasizes the disparity (Table 61.18). More refined rate estimates require a denominator that consists of the number of pesticide-exposed individuals. Unfortunately, such estimates are generally either imprecise or unavailable. For occupational exposures, the most straightforward denominator is the number of pesticideexposed individuals in the occupational group of interest (e.g., the number of licensed pesticide applicators). Regrettably, this information is unavailable. Little information is available on the magnitude of most pesticide-exposed workers, such as unlicensed applicators, or individuals who perform agricultural work. The transient nature of the agricultural workforce in many areas, and the fact that some agricultural workers may have undocumented U.S. immigrant status (i.e., lack of a U.S. visa or other immigration document), further complicates developing reliable denominators for pesticide-exposed agricultural workers. Considering variations in exposure among the agricultural workforce adds complexity to the problem (e.g., what pesticides are used, method of application, duration of use, and use of personal protective equipment). To calculate rates for nonoccupational pesticide-related illness, denominator information is needed on the use of pesticides by homeowners. This information is even more difficult to acquire than estimates of occupational pesticide users. Possible surrogates include data from a U.S. EPA 1990 survey of home and garden pesticide use that provided estimates of the number of containers and number of applications of pesticides for all households in the United States (Whitmore et al., 1992). Denominators can also be derived from pesticide use databases. Nationally, some survey data are available on the annual quantities of agricultural pesticides that were used (U.S. Department of Agriculture, 1998). Approximately every 2 years, the U.S. EPA publishes a report on pesticide sales and usage. The most recent report was published in 2004 and provides pesticide usage data through 2001 (Kiely et al., 2004). This document provides information on pesticide expenditures (in U.S. dollars) and amounts of pesticide active ingredients used (in pounds) both worldwide and in the United States. The data provided in that report are collected by various sources, including the U.S. EPA, the U.S. Department of Agriculture, and various private vendors. In addition, at least six states mandate some form of pesticide use reporting: Arizona, California, New Hampshire,

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TABLE 61.18 Annual Fatalities Ascribed to Pesticide Intoxication by Country Country (source of data)

Time period

Average annual % Self-inflicted no. of deaths (SD)

Approximate population of country (millions)

Average annual rate of all pesticide fatalities/million

Average annual rate of unintentional pesticide fatalities/milliona

Costa Rica (Wesseling et al., 1993)b

1980–1986

40.4

84

2

20.2

3.2

Jordan (Abdullat et al., 2006)

1999–2002

35

64

5.2

6.8

2.4

South Korea (Lee et al., 2009)

1996–2005

2536

85

48

53.2

7.9

Sri Lanka (de Alwis 1975–1983 and Salgado, 1988)c

1032 (200)

66

14.5

71.2

24.2

Sweden (Persson et al., 1997)

1969–1994

0.84

85

8

0.1

0.02

United Kingdom (Thompson et al., 1995)

1990–1991

22

66

50

0.4

0.01

United States (see Section 61.2.5)d

1997–2005

21

—

285

—

0.07

a

Excludes self-inflicted poisoning deaths. Data presented are crude figures derived from official sources and are believed more likely to be comparable to the data in the other entries in this table. A chart review conducted by Wesseling et al. (1993) identified additional cases (revised annual number of deaths per year  61.3) and called into question the percentage of deaths that were self-inflicted (revised estimate  62%). c Percentage self-inflicted based on review of a sample. d These data contain only unintentional cases. b

New York, New Jersey, and Oregon. The most comprehensive pesticide use reporting systems are found in California and New York. Beginning in 1996 in New York, commercial applicators must maintain records on all pesticide applications. The record for each application must include the U.S. EPA registration number, the name of the pesticide product, the amount applied, and the date and location of the application. Individuals who sell pesticides must maintain records on each pesticide purchase, including the U.S. EPA registration number, the name of the pesticide product, the date and amount purchased, and the intended location of the pesticide application. Information on all applications and pesticide purchases must be reported annually to the New York Department of Environmental Conservation. Detailed pesticide use data are available in New York through a written request to the New York Department of Environmental Conservation. Brief annual summaries of the data are available at the following website: http:// www.dec.ny.gov/chemical/27506.html. Since 1990 in California, growers, commercial pest control operators, and professional gardeners must submit

monthly reports on all pesticides that they use. The reported information must include the date and location of the application, the kind and amount of pesticide used, and the type of commodity if the pesticide was used on a crop. These data have been used to identify risk factors associated with illness from restricted-use organophosphate pesticides in an agricultural setting (Weinbaum et al., 1997) and were utilized to find that Parkinson’s disease mortality was higher in those counties where agricultural pesticides were used (Ritz and Yu, 2000). The CDPR collects and maintains the pesticide use data, but the Pesticides Action Network of North American website provides easy and convenient access to the data (see http://www.pesticideinfo.org/Search_Use.jsp). Problems have arisen with denominators derived from pesticide use databases. Among these are the lack of information on the number of exposed workers and their duration and intensity of exposure. In addition, care must be taken when using total poundage as the denominator. Pesticides applied at low rates may exhibit exaggerated risk if the time required for application is similar to or greater than the length of time to apply high-rate pesticides.

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When evaluating rates of pesticide-related illness, it is important to be mindful of the various factors that can influence the data. These include the methods used to approximate the person–time at risk, the inherent toxicity of the pesticide, the method of application, the amount applied, the equipment used, and the skills of the applicators. Any analyses that compare groups based on only some of these factors make the implicit assumption that the groups do not differ with respect to the other factors. Generalizing rates of poisoning to those outside the population that was investigated can be problematic. For example, the risks among fieldworkers can vary depending on the crops they tend, the tasks they perform, climatic conditions, and their individual work practices. The most difficult factors to ascertain may be the most critical. Ideally, changes in fieldworker tasks and practices over the time interval of interest should be determined. As such, when generalizing surveillance data on pesticide-related illness, the amounts of pesticide used or number of acres treated may not be adequate to assess the true risks that were experienced by fieldworkers.

61.6.2  Limitations Related to Definitions Enumerating the population at risk is not the only problem in interpreting surveillance data. When using surveillance data or comparing data from various surveillance systems, it is extremely important to understand the case definition that was used. Different pesticide poisoning surveillance systems may assign different meanings to the terms “pesticide,” “exposure,” “case,” and “related.” In the United States, FIFRA defines pesticides to include “all substances and mixtures of substances” that “prevent, destroy, repel, or mitigate” any pest (7USC§136). Other statutes clarify that pests include all deleterious organisms, even bacteria and viruses. Pesticides consequently include sanitizers and disinfectants (e.g., chlorine) along with moth balls, rat baits, herbicides, fumigants, and many other substances. Few programs attempt to track the effects of all these products. Most surveillance systems track illnesses and injuries associated with exposure to insecticides, fungicides, rodenticides, fumigants, and herbicides. Health effects from exposures to other pesticide products such as disinfectants and antibacterials are not universally included in surveillance systems. Each surveillance system sets its own threshold for the types of cases that are captured. Some systems attempt to identify only cases that resulted in medical consultation or only those that included hospitalization. Other data sources accept self-reports of illness or injury. Even when systems use the same objective standard, such as hospitalization, criteria for hospital admission may vary with culture, geography, time period, and economic circumstances. Additional discrepancies stem from variations in criteria for what constitutes pesticide-related illness and injury. Some have argued that surveillance should consider only

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those intoxications in which the pesticide acts on human victims by the same mechanism by which it controls pests (i.e., cholinesterase inhibition for organophosphate pesticides). Some critics have also taken issue with inclusion of health effects that may be related to the “inert” (nonpesticidal) ingredients in a pesticide product. Public health surveillance more typically considers all characteristics of pesticide products that can cause harm. These surveillance systems record illness and injury resulting from exposure to the pesticide products (i.e., the formulated product, not just the active pesticide ingredient). In addition, these surveillance systems record burns or smoke inhalation from pesticide fires, traumatic injuries from pesticide explosions, illnesses resulting from purposeful (i.e., homicidal or suicidal) and unintentional ingestion, allergic reactions to pesticide products, and effects associated with inert ingredients. They may even record reactions to a pesticide’s noxious odor. Evaluation of the causal relationship between pesticide exposure and adverse health effects is complicated. Some surveillance systems accept the clinician’s diagnosis in determining the relationship, whereas other systems have more complex case definitions and classification schemes. It is useful for surveillance systems to independently examine the relationship between exposure and health effects because the clinical diagnosis may not be correct (e.g., the health care professional may report illnesses suspected to be pesticide related but that are later found by surveillance staff to be unrelated to pesticide exposure). Surveillance systems that examine the relationship between exposure and health effects must take several factors into account, including the wide range of symptoms various pesticides can produce, the nonspecific nature of reported signs and symptoms (especially in less severe illness), limited or nonexistent analytical environmental data on the individual’s exposure, lack of clinical/biological measures of pesticide absorption, and inappropriate use of available tests. Evaluating anxiety poses a particular problem because many common insecticides are neurotoxic and may elicit anxiety through pharmacologic mechanisms; on the other hand, anxiety unaccompanied by pesticide exposure often mimics toxic effects. Rarely can physical findings or test results clarify this issue. In response to these difficulties, pesticide surveillance systems typically classify cases into one of several categories that reflect the certainty of the relationship between exposure and illness (see Section 61.5). When examining surveillance data, care must be taken not to confuse reports with confirmed cases. This is especially true for surveillance systems that include reports from affected individuals and nonmedical personnel and where no investigation is undertaken to follow up the report. Following an appropriate investigation, surveillance systems that use the NPHSS case definition classify reports of illness into one of the following categories: “confirmed” (i.e., defined as definite, probable, and possible cases),

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“nonconfirmed” (i.e., defined as unlikely cases, those determined not to be a case, and those where insufficient information is available), or “suspicious” (i.e., insufficient toxicologic information is available to determine whether a causal relationship exists between the exposure and the health effect). Some might argue that including case reports that are unconfirmed or have a lower degree of certainty compensates for underreporting. However, this may not be an appropriate remedy for underreporting because the unconfirmed cases may not be representative of the undetected true cases.

61.6.3  Limitations Related to Sensitivity No surveillance system succeeds in identifying every event of interest. Most surveillance systems capture 5–80% of cases that occur (Cates and Williamson, 1994). It should be recalled that even surveillance system data without high sensitivity can be useful for monitoring trends as long as the sensitivity remains relatively constant. The likelihood that a case of pesticide-related illness will be reported may vary with occupation, social status, the circumstances of exposure, and even the individual pesticide. Surveillance systems that rely on a variety of sources for case ascertainment are likely to be more representative of the universe of cases. Physician reporting, usually indirectly through workers’ compensation systems or poison control centers, is one of the most common mechanisms for surveillance. This method is the mainstay of many communicable disease reporting systems, but it is not necessarily the most effective method for surveillance of pesticide poisoning. Physician reporting requires that the affected individual seeks medical care, that a diagnosis of pesticide-related illness or injury is made or suspected, and that the physician reports the suspected case to public health authorities. Barriers exist at each of these steps that can hamper physician reporting. For example, some populations at greatest risk for pesticide exposure are likely to seek medical attention only for moderate or high severity illness. Those ill individuals who do not seek health care may be detected by surveillance systems based on physician reports only after an investigation into a sentinel case involving a poisoned co-worker or family member who sought medical care. Furthermore, pesticide-related illness is not routinely encountered by a majority of primary care providers in the United States, and most receive minimal training on recognition of environmental or occupational illness (Institute of Medicine, 1988; Pope and Rall, 1995; Schenk et al., 1996). In addition, the ability to make the diagnosis is complicated by the fact that symptoms are often nonspecific, and by the lack of readily available urine or blood tests to measure the pesticide, its metabolites, or the effects of the pesticide. Even when tests are available, they are frequently not performed, are used inappropriately (e.g., measuring

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cholinesterase depression after exposure to a pyrethroid insecticide), or are not performed sufficiently promptly to detect the abnormality. This problem of recognition is found in industrialized and developing countries alike (Keifer et al., 1996). Once the diagnosis of pesticide poisoning is made, there are many reasons for failing to report it. Despite broadly worded reporting guidelines, physicians are often reluctant to report cases that they believe are unconfirmed clinically. Additional barriers to physician reporting include protection of a patient who fears job loss, ignorance regarding the reporting requirement, and concerns that reporting a case may disrupt the personal relationship between the physician and the employer. Cultural pressures to downplay the hazards of pesticides and perceptions that registered pesticides are unlikely to cause illness may prevent a physician from reporting cases.

61.6.4  Legitimate uses for Surveillance Data With so many difficulties and limitations, surveillance of pesticide-related conditions may seem futile. However, even problematic surveillance data can advance public health when used with appropriate caution. Surveillance data have been useful for identifying emerging pesticide hazards and new populations at risk. When these emerging problems are identified, they present an opportunity to implement interventions that will prevent subsequent illness. For example, the identification of several California grape harvesters who became ill after exposure to phosalone led directly to the withdrawal of this pesticide. Although phosalone had been in use for nearly 20 years on crops that require minimal to moderate hand labor activity, it was eliminated only after it began to be used more widely on grapes, a crop requiring more extensive hand labor activity. This problem was detected when the ill grape harvesters were identified using surveillance data (O’Malley and McCurdy, 1990). A similar scenario was repeated in 1993 in Washington when 26 workers at 19 orchards became ill during a period of several months. The outbreak and the ensuing investigation resulted in the suspension, and eventual withdrawal, of mevinphos use in Washington apple and pear orchards (CDC, 1994; Washington State Department of Agriculture, 1994). Another example involves surveillance data from Florida that resulted in identification of illnesses associated with efforts to control Mediterranean fruit fly (medfly) infestations (CDC, 1999a). Medfly is an insect that can damage more than 250 fruit and vegetable plant species, and it is a serious threat to the agricultural industry. In the spring and summer of 1998, aerial applications of malathion combined with a corn protein bait were used by federal and state agriculture authorities to eradicate medfly infestations detected in portions of four Florida counties. A total

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of 123 acute pesticide-related illness cases associated with exposure to the pesticide used in the eradication effort were identified, representing a crude rate of 9 cases per 10,000 residents in the exposed areas. Following publication of these findings, the U.S. and Florida Departments of Agriculture adopted procedures to control medfly without the use of pesticides. These methods included more rapid detection of medfly infestations and the release of sterile male medflies to interrupt the reproductive cycle. To our knowledge, there have since been no medfly infestations in the United States that required aerial pesticide application for control. These situations exemplify the public health importance of prompt health care provider reporting and appropriate public health agency follow-up. Although surveillance that includes active case follow-up is resource intensive, it provides the opportunity to gather information that can be used to develop strategies for prevention. The information obtained through case follow-up is often more difficult to obtain or not available when cases are identified retrospectively through review of existing data sources (e.g., hospital discharge records). This is because long latency between a poisoning event and investigation makes locating the case more difficult, especially if the case is transient; can compromise the memory of events by those familiar with the poisoning incident; and can preclude an environmental investigation to identify residues or collateral damage because these will have vanished. Better surveillance likely would have reduced the health and financial costs associated with the indoor use of methyl parathion. From 1984 to at least 1997, homes and businesses in at least five different states were illegally sprayed with methyl parathion (Figure 61.10). Unfortunately, corrective action was not enacted until 1997. These events occurred in New York, Ohio, Michigan, Mississippi, and Illinois, resulting in expensive relocation and remediation activities. Relocations have involved more than 1500 individuals. The estimated cleanup costs for these incidents were more than $90 million (U.S. EPA, 1997a). Little information is available on the health effects associated with these incidents. However, one published report described methyl parathion-related illness among seven siblings, two of whom had a fatal outcome (CDC, 1984). In addition, another government report that summarized the 1995 Ohio investigations found that 20% or more of respondents reported symptoms during the 2 weeks following methyl parathion application (NCEH, 1996). These investigations also found that 20 of 50 (40%) indoor pet animals present in these homes died within 2 weeks of methyl parathion application (pointing out the value of pets as sentinels of human exposure and illness). To prevent additional exposure incidents, a memorandum of agreement between the U.S. EPA and the manufacturers of emulsifiable methyl parathion concentrate became effective in January 1997. The agreement included recall of particular products, changes in packaging and labeling,

Hayes’ Handbook of Pesticide Toxicology

FIGURE 61.10 Worker conducting methyl parathion remediation to a dwelling in Mississippi (courtesy of U.S. EPA).

as well as the addition of an odor-producing agent (U.S. EPA, 1997b). Because the problem with methyl parathion has existed since at least 1984 in several different states, a national surveillance system with active case follow-up may have resulted in a more timely identification of the magnitude of this problem and earlier adoption of preventive and regulatory measures. Surveillance data can also be a source of cases for formal epidemiologic studies. Important morbidity and mortality studies can be designed that involve a cohort of individuals poisoned by a specific pesticide or group of pesticides. For example, Steenland et al. (1994) examined a group of workers who had a history of acute organophosphate insecticide poisoning to determine if these workers had chronic neurologic sequela. The workers in this study were identified using surveillance data collected by the California EPA. As noted previously, surveillance data can also be used to examine the magnitude of pesticide-related illness and to assess trends. Data from several different surveillance systems have been combined to provide a more comprehensive estimate of the magnitude of pesticide-related illness and injury. For example, data from AAPCC, SENSORPesticides, and CDPR were combined to assess the magnitude and trends of pesticide poisoning among working youth (Calvert et al., 2003), the adverse effects of automatic insecticide dispensers (CDC, 2000), and illnesses associated with pesticide exposures at schools (Alarcon et al., 2005).

61.6.5 Mechanisms to Strengthen the Surveillance of Acute Pesticide-Related Illness Some very specific actions can be taken to enhance existing surveillance systems. Some of the changes have been underway for several years, and an evaluation of their efficacy should be possible.

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Figure 61.11  A scout uses the “white pan beat method” to check strawberries for insects (courtesy of California EPA).

One important action is the need to improve training of primary health care professionals. The U.S. EPA launched an initiative to target health care professionals with educational and training opportunities on pesticide-related health issues (Lindell et al., 2003a,b; National Environmental Education and Training Foundation, 2002). This initiative involves strategies to ensure that primary health care professionals can recognize health effects from pesticide exposure. It includes mechanisms to enhance their abilities to diagnose illness and manage exposures, engage in preventive management (at the case and community level), appropriately report exposures and illnesses to public health authorities, and access appropriate resources when necessary. To foster success, these activities should be coupled with education of workers and consumers on many of these same topics. The public should be warned about pesticide dangers through broad media campaigns that explain the importance of reading and understanding the pesticide label, using pesticides sparingly in conjunction with implementation of an integrated pest management plan (Blessing, 2001), and taking necessary precautions (e.g., using protective equipment such as chemical-resistant gloves) (Figure 61.11). Also, pesticide labels should be improved to make information easier to find and understand. Increasing the quality and availability of biological monitoring tools would aid surveillance by assisting with

confirmation of cases. The development of new biomarkers of exposure and health effects is also an extremely important area that would enhance surveillance data. Reliable and affordable screening methods for field and clinical settings must be available if they are to be used routinely in developing countries and under the constraints of managed health care systems. Most pesticide biological markers are still primarily research tools, and analyses are conducted only at specialized centers (e.g., CDC Environmental Health Laboratory and NCEH). Even cholinesterase monitoring, the most commonly used measure of biological effects from exposure to organophosphate and carbamate insecticides, suffers from lack of standardization. Both the handling of specimens and the assay method require standardization to obtain valid test results (Wilson et al.,1996, 2004). Although much progress has been made in delineating these problems, they have not been satisfactorily resolved (Wilson et al., 2005). The ability to provide summary data and direct feedback to the medical community, agricultural workers, pesticide manufacturers, commercial pest control firms, and policy makers is a critical aspect of surveillance. Although existing surveillance systems have communicated their findings in several publications, more can be done to share surveillance findings with various stakeholders, including partner government agencies, public interest groups (e.g., environmental groups and public health organizations),

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agricultural employers, pesticide manufacturers, the pest control industry, and worker advocacy groups (e.g., unions). In addition, the ability to aggregate data across states, combined with increased dissemination of information, has resulted in a better understanding of the extent and nature of acute pesticide-related illness and injury. However, this is also an area in which improvements can be made that will strengthen surveillance. If a pesticide poisoning surveillance program does not have a statutorily mandated multiagency oversight committee (or board) to address pesticide use and hazards, the program might benefit from developing an advisory committee. Members can include stakeholder representatives as described in the previous paragraph. Advisory committee meetings, which should occur two to four times per year, are often a source of valuable ideas for the program. They also provide the pesticide poisoning surveillance program an opportunity to maintain contact with various constituencies, apprize them of findings, develop joint programs for outreach and intervention, and discuss mechanisms for improving reporting and investigation. Lack of access to health care may be an important barrier to identifying acute pesticide-related illness among individuals with low incomes, including agricultural workers and their families. The most obvious solution would be to improve access to health care. Because farm workers are at greatest risk of acute pesticide poisoning (Calvert et al., 2008), a periodic survey of the farm worker population would be useful to assess the magnitude and trend of pesticide poisoning in this population. One approach would be to piggy-back onto NAWS (see Section 61.2.9). In 1999, NAWS included questions to determine the magnitude of pesticide poisoning among crop workers. To our knowledge, neither these nor similar questions to assess the incidence of pesticide poisoning were included in NAWS surveys before or after 1999.

61.7  Fundamentals of epidemiology When conducting surveillance of pesticide-related disease and injury, a decision must be made as to whether the pesticide exposure caused the documented illness. Epidemiologic studies are often the source of information used to make these decisions. Therefore, although the emphasis of this chapter is on surveillance, we think it is important to describe the basic principles of epidemiology and the role of epidemiology for identifying health effects related to pesticide exposure. It is useful to begin by defining epidemiology. Epidemiology is the study of the distribution and determinants of disease in human populations. Epidemiologic studies compare the rates of disease in populations exposed to various risk factors to the rates in populations that are not exposed. Such comparisons evaluate what factors may cause or influence disease.

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61.7.1  Principles of Epidemiology A major premise of epidemiology is that disease is not simply a random occurrence but, rather, is the result of various causal factors (Checkoway et al., 1989). These causal or risk factors influence the distribution of disease in a population. Differences in disease patterns may be explained by differential distribution of risk factors between populations. These causal or risk factors include age; sex; race or ethnicity; genetic susceptibility; personal lifestyle factors such as smoking habit, diet, exercise, use of drugs, and weight; and occupational or environmental exposure to various chemical and physical agents. Diseases caused by occupational or environmental exposures may be either acute or chronic. Acute diseases occur soon after an exposure, whereas chronic diseases develop many years after exposure. Examples of acute disease are leukopenia caused by acute radiation poisoning and eye and upper respiratory irritation caused by sulfur dioxide or ozone. Examples of chronic disease include pneumoconiosis caused by crystalline silica or coal dust and asthma caused by isocyanates. The time period between initial exposure to a causal agent and disease detection can be divided between the induction period – time between causal action and disease initiation – and latency period – time between disease initiation and detection. The longer the induction–latent period, the more difficult it is to link causal factors to the disease outcome. Pesticides are known to cause both acute diseases, such as systemic poisonings and skin rashes, and chronic diseases, such as cancer, lung disease, neurotoxicity, and reproductive problems (Alavanja et al., 2004).

61.7.2  Epidemiologic Study Designs When studying the health effects of occupational and environmental exposures, epidemiologists are very rarely able to control the exposures administered to study subjects, such as in a randomized controlled trial of a pharmaceutical. Therefore, epidemiology of occupational and environmental exposures is largely an observational science, relegated to documenting the past and present risk factors and evaluating the association between these risk factors and disease status (Kleinbaum et al., 1982). Epidemiological research begins with a hypothesis to be tested. For example, it may be hypothesized that farmers with a history of applying insecticides to their crops have a greater risk of developing cancer than the general population. The epidemiologist then designs a study to test this hypothesis. The epidemiologist may choose from among the following types of observational (nonrandomized) study designs. For a fuller discussion on these study designs and their interpretations, the reader is referred to standard epidemiologic textbooks (Kleinbaum et al., 1982; Rothman et al., 2008).

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(a) Cohort Study Cohort studies begin with enumeration of a population (the cohort) that shares common characteristics or risk factors (e.g., exposures). The cohort’s health experience is then evaluated over a defined time period. The following is the basic question addressed by a cohort study: Are those with exposure more (or less) likely to develop disease compared to those who are unexposed? The control or reference populations are generally national or regional (i.e., state or province) populations, which provide generally stable disease rates. Cohorts can be enumerated in present time and then followed forward in time. This approach is termed a prospective cohort study. As the cohort is followed through time, the exposures and diseases of the individual cohort members are documented. The rates of disease among individuals with specific exposures of interest are compared to the rates of disease among unexposed or low exposed members or an unexposed reference population. Another type of cohort study is the retrospective cohort study, which is also referred to as a historical cohort study. In this type of study, a cohort is enumerated in the past (i.e., a cohort that began exposure at some point in the past) and followed up to the present to identify those individuals who developed disease. The disease rates in the cohort are compared to those occurring in an unexposed comparison population. An example of a prospective cohort study is the Agricultural Health Study (AHS), which was begun in 1994. The AHS is being conducted in Iowa and North Carolina by the National Cancer Institute, the National Institute for Environmental Health Sciences (NIEHS), the U.S. EPA, and NIOSH (Alavanja et al., 1996). It consists of 57,311 licensed restricted-use pesticide applicators and 32,347 of their spouses. The AHS participants are asked to answer questionnaires about their lifestyle, work practices, and exposures at approximately 5-year intervals in an attempt to relate exposure to disease outcomes. The participants are also matched to vital records systems and cancer registries to identify causes of death and cancer incidence, respectively. Disease rates are determined at regular intervals, and potential risk factors are assessed by the information collected on the questionnaires and contained in disease registry records. An internal reference population is usually used that consists of the unexposed or low-exposed portion of the AHS cohort. This study has been prolific in generating important findings, including associations between carbofuran and lung cancer (Bonner et al., 2005), alachlor and lymphohematopoietic cancers (Lee et al., 2004), and diazinon and lung caner and leukemia (Freeman et al., 2005). An example of a retrospective cohort study is provided by a study of 20,245 agricultural pesticide applicators in Sweden (Wiklund et al., 1989). It also evaluated the association between pesticide exposures and disease.

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The cohort, consisting of all applicators who had been licensed between 1965 and 1976, was followed in the Swedish Cancer Register from the date they received their license until 1982 or their death if it occurred before 1982. The number of cancer cases in the applicator cohort was compared to the number of cases occurring in 5-year age and sex groups during the same time period in the whole Swedish population. The age-, sex-, and time periodadjusted number of cancer cases in the Swedish population represent the number of cases that would be expected in the pesticide applicator population. A total of 558 malignant tumors were found among Swedish pesticide applicators compared with 649.8 expected cases, resulting in a statistically significantly decreased standardized incidence ratio of 0.86. A finding that reaches statistical significance implies that the finding is unlikely to be due to chance. No cancer rates among the Swedish pesticide applicators were found to be significantly increased, although they had higher rates for testicular cancer, Hodgkin’s disease, and tumors of the nervous system and endocrine glands. Cohort studies have several strengths. They provide information on the time lag between the first known exposure and disease detection, and they can be used to evaluate risk for many different diseases. They also measure exposure before disease occurs, resulting in less recall bias by study subjects. However, cohort studies are costly, requiring long-term commitments of time and resources and a large sample size. In addition, they are ill-suited for studying rare diseases. Furthermore, a cohort study may not be possible if data for enumerating and following up the cohort are incomplete. Retrospective cohort studies are usually restricted to investigating fatal diseases because nonfatal diseases (with the exception of cancer and endstage renal disease) are often not recorded historically. In contrast, prospective cohort studies may document nonfatal diseases as they occur. The primary cost for conducting cohort studies often involves obtaining exposure data on a large number of subjects, of which only a small proportion develop the disease of interest. (b) Case–Control Study The following is the basic question addressed by the case– control study: Are the people with existing disease more or less likely to have been exposed than those without the disease? The distinguishing feature of case-control studies is that subjects are selected based on their disease status, reducing cost by limiting exposure assessment to only cases of the disease of interest and a control group. Case– control studies are particularly useful for studying rare diseases or diseases with a long latency since first exposure. Cases may be identified from disease registries, hospital or clinical records, or volunteers. Controls are selected to be similar to the cases with exception of disease status and are often matched to cases by age, gender, race, and residence.

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Exposure information is obtained either from existing records or from a detailed self-reported questionnaire. This information is used to compare the exposure prevalence between the cases and controls. An example of a case–control study to evaluate the association between pesticide exposure and disease is provided by a study of childhood brain cancer in four Atlantic coast states (Shim et al., 2009). All cases of brain cancer were diagnosed when the children were younger than 10 years of age, between the years 1993 and 1997, and were identified from statewide cancer registries in four states (i.e., Florida, New Jersey, New York, and Pennsylvania). Each case was matched to one control through telephone calls using random-digit dialing (Waksberg, 1978). Controls were individually matched to cases by age (within 1 year of the birth year), sex, race, and state of residence. The mothers of the cases and controls were interviewed via a computer-assisted telephone system over 13 months between 2000 and 2001. The questionnaire obtained, among other things, information on maternal and paternal residential and occupational pesticide exposures during the 2 years before the child’s birth. Analyses were separately performed for astrocytoma and primitive neuroectodermal tumors (PNET), the two most common types of childhood brain cancer. Residential herbicide use was associated with a significantly increased risk for astrocytoma (odds ratio (OR), 1.9; 95% confidence interval, 1.2–3.0), as was residential and occupational use of herbicides by the father (OR, 1.8; 95%CI, 1.1–3.1). PNET was not found to be associated with pesticide exposure. Case–control studies are relatively inexpensive because they involve fewer subjects than cohort studies and can be completed in a relatively short time. However, the information collected on exposures occurs after the disease has been diagnosed, which may make diseased people more (or less) likely to remember previous exposures (recall bias). Also, diseased individuals may be more motivated to participate in a case–control study than a healthy control (selection bias). (c) Cross-sectional Study In a cross-sectional study, exposure and disease are evaluated at the same time. The prevalence of disease is measured in a defined population at a particular point in time, whereas the exposures of the individuals are also measured at that time and retrospectively. For example, the rate of occurrence of symptoms in workers exposed to a particular pesticide would be compared to the rate in workers who were unexposed. Cross-sectional studies are suitable for evaluating nonfatal diseases or measuring physiologic responses to workplace exposures. Data are collected using clinical examinations, symptom surveys, or direct biological or physical measurements. A critical problem with the cross-sectional design is the “chicken or the egg” conundrum – that is, did the exposure influence the disease or

Hayes’ Handbook of Pesticide Toxicology

vice versa. Also, cross-sectional studies may miss many diseases of short duration. Because cross-sectional studies typically only include currently employed workers, retirees or other workers who terminated employment because of ill health, possibly attributable to their exposures, are not studied. Exclusion of these former workers may bias the study findings toward the null because these former workers may be the most relevant subjects for investigating delayed or progressive health outcomes. An example of a cross-sectional study is one that was used to evaluate the association between fumigant exposure and disease among structural fumigation workers in Florida (Calvert et al., 1998). In this study, 123 structural fumigation workers and 120 unexposed controls were interviewed and examined. Nerve conduction, vibration, neurobehavioral, olfactory, visual, and renal function testing was conducted. The median lifetime duration of methyl bromide and sulfuryl fluoride exposure among workers was 1.20 and 2.85 years, respectively. Sulfuryl fluoride exposure over the year preceding examination was found to be associated with significantly reduced performance on one cognitive test and on olfactory testing. In addition, fumigation workers had significantly reduced dexterity of the dominant hand. A nonsignificantly higher prevalence of carpal tunnel syndrome was also observed among the fumigation workers. The authors concluded that occupational sulfuryl fluoride exposures may be associated with subclinical effects on the central nervous system, including effects on olfactory and some cognitive functions. However, no widespread pattern of cognitive deficits was observed. The peripheral nerve effects were likely due to ergonomic stresses experienced by the fumigation workers. (d) Ecologic Study Perhaps the crudest approach to evaluating exposure– disease associations is the ecologic study. In this type of study, disease rates are compared between geographic areas rated according to their estimated extent of exposure. The units of exposure correspond to geographical areas rather than individuals. An example of an ecologic study is one used to evaluate the association between cancer and dibromochlorpropane (DBCP) contamination in Fresno County, California, drinking water (Wong et al., 1989). All cases of gastric cancer and leukemia cases occurring between 1960 and 1983 in Fresno County were identified by the California Vital Statistics office. The cancer rates were calculated using the 1960, 1970, and 1980 census data stratified by age, sex, and race. The cancer rates were compared by areas in the county stratified by the concentration of DBCP found in drinking water. No correlation was found between gastric cancer and leukemia mortality rates and DBCP concentrations. Ecologic studies use readily available data that have been collected for other purposes, and they can be done

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relatively quickly. However, ecologic studies are severely limited because they do not associate exposure to individuals. They suffer from the “ecological bias” (Wakefield, 2008) that states that conclusions regarding individual risk on the basis of group risk must be made cautiously because risk factors at the individual level have not been collected. Ecologic studies do not control for other exposures (confounding) that may be associated with the disease of interest, and they can be significantly influenced by the migration of individuals into or out of the geographic area (selective migration). Although ecologic studies do not provide firm conclusions about the association between exposures and disease, they are used for generating hypotheses and to guide future, more in-depth research. (e) Case Reports and Case Series These consist of a report of a medical diagnosis among an individual or series of individuals with exposures not previously thought to be associated with that disease. Sometimes these are referred to as disease clusters. These can be particularly informative when the disease is rare and when etiologic factors are unknown. Like cross-sectional studies, case reports and case series provide little information on cause and effect. However, they are important sentinels that alert epidemiologists of suspected disease– exposure relationships that require more in-depth epidemiologic investigation. An example of a case report is an account of a birth defects cluster involving three infants born within 8 weeks of each other to pesticide-exposed migrant farm workers employed in North Carolina and Florida (Calvert et al., 2007a). During the period of organogenesis (approximately days 14–59 after fertilization) when birth defects are most likely to occur, all three mothers appear to have unknowingly worked in tomato fields that were under a restricted entry interval because the fields were recently treated with pesticides, some of which have been shown to be animal teratogens. One infant was born with tetraamelia (the absence of all four limbs). The second infant was born with mild Pierre Robin syndrome (micrognathia, high arched palate, and mild persistent palatine rugae). The third infant died at 3 days of age with multiple severe malformations, including cleft lip and palate, imperforate anus, solitary kidney, vertebral anomalies, dysplastic lowset ears, and ambiguous genitalia. These findings were reminiscent of a severe type of the Goldenhar syndrome (also referred to as oculoauriculovertebral sequence). The cluster was investigated by NIOSH and public health authorities in Florida and North Carolina. It was concluded that the available evidence was inadequate to establish a causal relationship between the birth defects and pesticide exposures. However, this cluster pointed to the need to better protect farm workers from pesticide exposures. In addition, following the cluster investigation, lawmakers in North Carolina enacted legislation to broaden the coverage

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of antiretaliation rules to include agricultural workers; increased recordkeeping requirements pertaining to pesticide applications; and funded various activities to prevent harm from pesticides, including strengthening surveillance, improving the quality of pesticide compliance inspections, and increasing and improving pesticide safety training (Calvert and Higgins, 2009). Disease clusters can be misleading. Occurrences of rare disease are expected to be distributed randomly. Random distributions can include clusters of disease. As such, one needs to remember that case series or disease clusters can occur randomly, and such clusters may not be indicative of an association with a hazardous exposure.

61.7.3  Evaluating Pesticide Health Information Information on the toxicity of pesticides can be found in textbooks, journal articles, and from information provided by pesticide producers. Each of these sources may be valuable, but each may also have particular bias. When relevant information is identified, one must decide whether to trust the information and whether the information can be generalized to new cases identified through surveillance. In many medical disciplines, information relevant to a particular question may be summarized in systematic reviews that provide valid conclusions. Unfortunately, such reviews are uncommon with respect to the health effects associated with specific pesticide products or active ingredients. As such, skill is needed both to efficiently search the literature for relevant information and to interpret the validity of any information that is discovered. There are several approaches for identifying relevant literature. These include asking knowledgeable colleagues (and pesticide producers), reviewing references cited in textbooks, or using an electronic bibliographic database such as PubMed. The limitation of asking colleagues and using textbooks is that these sources may not be up-to-date. The most up-to-date source of relevant information is obtained by searching PubMed. Accessing this database has become a basic and easily acquired skill. There is no charge for accessing this database through the National Library of Medicine (http://www.ncbi.nlm.nih.gov/pubmed). Assessing the validity of a study can be more complicated (Levine et al., 1994). In cohort and cross-sectional studies, it is important that the exposed and control groups are similar with respect to all factors that may affect the outcome, with the exception that the exposed group has the exposure of interest. Practically, this means that both groups should be similar in age, gender, race, socioeconomic status, cultural background, and social habits (e.g., smoking and alcohol consumption). The study investigators should demonstrate that these characteristics are comparable or use statistical techniques to adjust for

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differences. In a case–control study, it is important that the cases with the disease of interest and the unaffected controls are similar with respect to important determinants of the disease of interest. In addition, controls should be randomly drawn from the same population from which cases were drawn (i.e., controls should have similar opportunity for exposure). For example, in a case–control study examining the association between pesticide exposure and cancer, it would be inappropriate to have the control group consist entirely of white-collar professionals because of their minimal opportunity for pesticide exposure. Even when investigators take appropriate precaution to ensure comparability for known risk factors, there may be a pronounced imbalance in the distribution of risk factors that are unknown to the investigators. It may be these unknown and unmeasured risk factors that are responsible for any observed findings. Other factors to consider when assessing the conclusions of an observational study include assessing the temporal sequence of events. It is important that the investigators document that the exposure preceded the disease or outcome of interest. It is also important to observe a dose– response relationship. Attributing a particular outcome to a pesticide exposure can be made with more confidence if risk of the outcome increases with increasing cumulative, intensity, or duration of exposure. It is also important to ensure that the investigators minimized bias in the collection of data (e.g., blinding data collectors and study subjects to the study hypotheses). The magnitude of the risk can also be helpful in assessing the validity of a study. With very large values of risk, one may be more confident that bias or uncontrolled risk factors are not responsible for the outcome. Note that none of the conditions described in this section are either necessary (with the exception of temporal sequence) or sufficient to prove that a causal association exists between an exposure and an illness. The topic of causality is complex and is not discussed further. An excellent review of this topic can be found elsewhere (Rothman et al., 2008).

61.8  Internet and telephone resources for pesticide information Important sources of pesticide information are readily available on the Internet. Some Internet sites that are useful for pesticide-related illness and injury surveillance follow.

(a) Agricultural Health Study This is a large prospective cohort study consisting of 57,311 licensed restricted-use pesticide applicators and

32,347 of their spouses. The AHS is being conducted in Iowa and North Carolina by the National Cancer Institute, NIEHS, the U.S. EPA, and NIOSH. This study is an important source of information on illnesses and their association with pesticide exposures. The website has information on the study design, important findings from the study, and guidelines for initiating research collaboration with the AHS (http://aghealth.nci.nih.gov/index.html). (b) Bureau of Labor Statistics This site contains pesticide illness tabulations that supplement those provided in the BLS annual reports. Information is available only for pesticide illnesses that result in lost worktime. The website address is http://www. bls.gov/iif. (c) California Environmental Protection Agency This site is a source for consumer fact sheets and provides access to several useful databases (e.g., a pesticide use in California database, a pesticide product database, a chemical ingredient database, a company information database, and a California pesticide illness query). The site also has links to the Pesticide Illness Surveillance Program. The website address is http://www.cdpr.ca.gov. (d) Extoxnet This is an excellent resource for information on specific pesticides. In includes a useful search engine. This site is a cooperative effort of the University of California–Davis, Oregon State University, Michigan State University, Cornell University, and the University of Idaho. The website address is http://ace.orst.edu/info/extoxnet/ghindex. html. (e) National Agricultural Statistics Service, U.S. Department of Agriculture This site provides information on agricultural pesticide usage. The website address is http://www.nass.usda.gov. (f) National Institute for Occupational Safety and Health NIOSH is the federal agency responsible for conducting research on occupational disease and injury. Its pesticide illness and injury surveillance topic page contains a wealth of useful information, including a guide on initiating and maintaining a state-based pesticide poisoning surveillance program (http://www.cdc.gov/niosh/docs/2006-102). This site also has information on the SENSOR-Pesticides program, including its case definition, standardized variables, fact sheets, a pesticide illness database, and summaries of surveillance findings. Its pesticide illness and injury surveillance topic page can be found at http://www.cdc. gov/niosh/topics/pesticides.

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(g) National Pesticide Information Center This is a cooperative effort of Oregon State University and the U.S. EPA. It offers a toll-free telephone service that provides information about pesticides and pesticide-related topics (1-800-858-7378). At its website, one can find fact sheets, information on current pesticide topics, podcasts, and web links to various pesticide-related topics (http:// npic.orst.edu). (h) Pesticides and Epidemiology: Unraveling Disease Patterns This is a useful primer intended to facilitate the understanding of basic epidemiological concepts and methods. Such an understanding will enable appreciation of scientific reports that explore associations between diseases and pesticide exposures. This primer is available at http://www. btny.purdue.edu/Pubs/PPP/PPP-43.pdf. (i) PubMed The medical literature can be searched for information on specific pesticides. Searches on this site are free-of-charge. The website for PubMed is http://www.ncbi.nlm.nih. gov/pubmed. (j) University of Nebraska–Lincoln Institute of Agriculture and Natural Resources This website contains an extensive listing of links on pesticide-related topics. The links are grouped into several categories, including environmental protection, pesticide licensing, pesticide labels and MSDS, pest management, and health protection and safety. The website for these pesticide-related links is http://pested.unl.edu/pesticide/ pages/index.jsp. (k) U.S. Environmental Protection Agency This site contains a large amount of information, including consumer fact sheets, information on pesticide regulations, and pesticide product information. The website is http:// www.epa.gov/pesticides. Pesticide product information can be found at http://www.epa.gov/opppmsd1/PPISdata/index. html. The manual titled “Recognition and Management of Pesticide Poisonings” is available at http://www.epa.gov/ oppfead1/safety/healthcare/handbook/handbook.htm.   An update of this manual is in preparation.

Conclusion A comprehensive, national surveillance system for acute pesticide-related illness and injury does not currently exist. However, this chapter describes several surveillance systems for pesticide-related illness and injury, each having strengths and weaknesses. Some systems, such as the NPDS, are most useful for assessing magnitude and trends.

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Others (e.g., state-based surveillance systems) are more useful for timely identification of outbreaks and emerging problems. Efforts to standardize data collection have been ongoing. Standardization of data collection facilitates linkage of data across surveillance systems to create a fuller understanding of the acute pesticide-related illness and injury problem. Through standardization and information sharing across surveillance systems, a national comprehensive surveillance system may be attainable. Although all of the surveillance systems described in this chapter have strengths and weaknesses, almost all provide useful information that is vital to target scarce public health resources toward the prevention of pesticide-related effects on human health and the environment.

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Schnitzer, P. G., and Shannon, J. (1999). Development of a surveillance program for occupational pesticide poisoning: Lessons learned and future directions. Public Health Rep. 114, 242–248. Shim, Y. K., Mlynarek, S. P., and van Wijngaarden, E. (2009). Parental exposure to pesticides and childhood brain cancer: U.S. Atlantic Coast Childhood Brain Cancer Study. Environ. Health Perspect. 117, 1002–1006. Simpson, W., Schuman, S., Caldwell, S., and Spell, L. (2004). Hospitalizations and emergency room visits for acute pesticide poisonings in South Carolina, 1997–2001. JSC Med. Assoc. 100, 39–42. Stanbury, M., Anderson, H., Rogers, P., Bonauto, D., Davis, L., Materna, B., and Rosenman, K. (2008). “Guidelines for Minimum and Comprehensive State-Based Public Health Activities in Occupational Safety and Health,” DHHS Publication No. 2008-148 Available at http://www.cdc.gov/niosh/docs/2008-148 (accessed June 1, 2009). National Institute for Occupational Safety and Health, Cincinnati, OH. Steenland, K., Jenkins, B., Ames, R. G., O’Malley, M., Chrislip, D., and Russo, J. (1994). Chronic neurological sequelae to organophosphate pesticide poisoning. Am. J. Public Health 84, 731–736. Thacker, S. B., and Berkelman, R. L. (1988). Public health surveillance in the United States. Epidemiol. Rev. 10, 164–190. Thompson, J. P., Casey, P. B., and Vale, J. A. (1995). Deaths from pesticide poisoning in England and Wales, 1990–1991. Hum. Exp. Toxicol. 14, 437–445. Thundiyil, J. G., Stober, J., Besbelli, N., and Pronczuk, J. (2008). Acute pesticide poisoning: A proposed classification tool. Bull. World Health Organ, 86(3), 205–209. Available at http://www.who.int/bulletin/volumes/86/3/07-041814.pdf (accessed June 15, 2009). U.S. Bureau of the Census (1992). “1990 Census of Population and Housing. Alphabetical Index of Industries and Occupations.” U.S. Department of Commerce, Bureau of the Census, Washington, DC. U.S. Census Bureau (2006). “Statistical Abstract of the United States: 2006,” Available at http://www.census.gov/prod/www/abs/ statab2001_2005.html (accessed May 6, 2009). U.S. Census Bureau, Washington DC. U.S. Department of Agriculture (USDA) (1998). “Agricultural Chemical Usage. 1997 Restricted Use Pesticides Summary Reports.” USDA, Washington, DC. Available at http://usda.mannlib.cornell.edu/ MannUsda/viewDocumentinfo.do?documentID1001. (Accessed October 20, 2009.)��������� U.S. Environmental Protection Agency (EPA) (1992). “Regulatory Impact Analysis of Worker Protection Standard for Agricultural Pesticides.” Available at http://www.aei-brookings.org/admin/authorpdfs/page. php?id821 (accessed June 15, 2009). U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (EPA) (1997a). “Interim Guidance on Maximizing Insurers’ Contributions to Responses at Residences Contaminated with Methyl Parathion.” Memorandum from Barry Breen, Director, Office of Site Remediation Enforcement. U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (EPA) (1997b). “Illegal Indoor Use of Methyl Parathion,”. Office of Pesticide Programs, U.S. EPA, Washington, DC. Available at http://www.epa.gov/pesticides/ factsheets/chemicals/methyl.htm. (Accessed ������������������������������������ October 20, 2009.)�������� U.S. Government Accountability Office (2000). “Pesticides: Improvements Needed to Ensure the Safety of Farmworkers and Their Children,” GAO/RCED-00-40. Available at http://www.gao.gov/new.items/ rc00040.pdf (accessed June 16, 2009). U.S. General Accounting Office, Washington, DC.

Chapter | 61  Surveillance of Pesticide-Related Illness and Injury in Humans

Vorhaus, L. J., and Kark, R. M. (1953). Serum cholinesterase in health and disease. Am. J. Med. 14, 707–719. Wakefield, J. (2008). Ecologic studies revisited. Annu. Rev. Public Health 29, 75–90. Waksberg, J. (1978). Sampling methods for random digit dialing. J. Am. Stat. Assoc. 73, 40–46. Washington State Department of Agriculture (1994). “Rules Restricting the Use of Phosdrin Finalized,” Press Release April 18, 1994. Washington State Department of Agriculture, Olympia. Weinbaum, Z., Schenker, M. B., Gold, E. B., Samuels, S. J., and O’Malley, M. A. (1997). Risk factors for systemic illnesses following agricultural exposures to restricted organophosphates in California, 1984–1988. Am. J. Ind. Med. 31, 572–579. Wesseling, C., Castillo, L., and Elinder, C. (1993). Pesticide poisonings in Costa Rica. Scand. J. Work Environ. Health 19, 227–235. Whitmore, R.W., Kelly, J.E., Reading, P.L. (1992). National Home and Garden Pesticide Use Survey Final Report, RTI/5100/17-01F. Research Triangle Institute, Research Triangle Park, NC. Wiklund, K., Dich, J., Holm, L. E., and Eklund, G. (1989). Risk of cancer in pesticide applicators in Swedish agriculture. Br. J. Ind. Med. 46, 809–814. Wilson, B. W., Padilla, S., Henderson, J. D., Brimijoin, S., Dass, P. D., Elliot, G., Jaeger, B., Lanz, D., Pearson, R., and Spies, R. (1996).

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Factors in standardizing automated cholinesterase assays. J. Toxicol. Environ. Health 48, 187–195. Wilson, B. W., Henderson, J. D., Arrieta, D. E., and O’Malley, M. A. (2004). Meeting requirements of the California cholinesterase monitoring program. Int. J. Toxicol. 23, 97–100. Wilson, B. W., Arrieta, D. E., and Henderson, J. D. (2005). Monitoring cholinesterases to detect pesticide exposure. Chem. Biol. Interact. 15, 157–158, 253–256. Wong, O., Morgan, R. W., Whorton, M. D., Gordon, N., and Kheifets, L. (1989). Ecological analyses and case–control studies of gastric cancer and leukemia in relation to DBCP in drinking water in Fresno County, California. Br. J. Ind. Med. 46, 521–528. World Health Organization (WHO) (1977). “Manual of the International Statistical Classification of Diseases, Injuries and Causes of Death,” 9th rev. ed. WHO, Geneva. World Health Organization (WHO) (1992). “International Statistical Classification of Diseases and Related Health Problems,” 10th rev. ed. WHO, Geneva. World Health Organization (WHO). (2002). “Report of the First Pesticide Surveillance Group Meeting,” Washington, 5–6 July 2002 [Internal report]. WHO, Geneva.

Chapter 62

Risk Assessment and Risk Management: The Regulatory Process Penelope A. Fenner-Crisp U.S. Environmental Protection Agency (Retired)

62.1  Introduction In the United States, primary authority for pesticide regulation resides with the U.S. Environmental Protection Agency (EPA) under the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA) and the Federal Food, Drug, and Cosmetic Act (FFDCA). FIFRA authorizes EPA to prescribe the con­ ditions of use of pesticide products. Under FFDCA, EPA establishes maximum allowable levels of pesticide residues (“tolerances”) in foods and animal feeds. The tolerances for most raw and processed foods are enforced by the Food and Drug Administration (FDA) of the Department of Health and Human Services (HHS) The U.S. Department of Agriculture (USDA) is responsible for enforcement of tolerances for meat, poultry, and some egg products.

62.2  Historical background of pesticide regulation in the united states Regulation of pesticides at the federal level has been in place for nearly a century. Each time legislation has been amended, the number and nature of the directives have been increased and embellished. Some, but not all, of these changes will be described. Only those areas of the law(s) that remain in effect under current legislation are discussed. The first relevant legislation passed was the Federal Insecticide Act of 1910 (FIA, 1910). The provisions of this act, essentially only a labeling statute, were limited to the prohibition of the manufacture of any insecticide or fungi­ cide that was “adulterated or misbranded.” No requirement for registration and no establishment of standards of safety were included at that time. Congress passed the first version of the Federal Insecticide, Fungicide, and Rodenticide Act in 1947, adding the requirement of registration by the Secre­ tary of Agriculture before sale or distribution in interstate or Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

foreign commerce, but without providing the U.S. Department of Agriculture (USDA) the power to deny or cancel a regis­ tration if the registration did not comply with the provisions of the law (FIFRA, 1947). A 1964 amendment to FIFRA did provide USDA with the authority to deny or rescind a regis­ tration and to issue an immediate suspension of registration if necessary to prevent an imminent hazard to human health (FIFRA, 1964). Nearly half a century after Congress passed the first federal pesticide regulatory legislation, it amended the Federal Food, Drug, and Cosmetic Act to require the Food and Drug Administration to establish maximum acceptable levels (“tolerances”) for pesticide residues in foods and ani­ mal feeds (FFDCA, 1954). This requirement applied only to raw agricultural commodities. Four years later, Congress once again amended FFDCA to include a requirement for a tolerance in a processed food, but only if the pesticide residue in that processed food was expected to exceed the tolerance level in the related raw agricultural commodity (FFDCA, 1958). In 1970, under President Nixon’s government Reorga­ nization Plan No. 3, the primary federal authority for the regulation of pesticides was transferred from USDA and FDA to the newly-created EPA (Nixon, 1970). Between 1970 and 1990, Congress amended FIFRA six times, each time adding to, or enhancing, the Agency’s exist­ ing responsibilities. One of the provisions added in 1972 was that a pesticide could be registered only if it did not cause “unreasonable adverse effects” on human health or the environment [Federal Environmental Pesticide Control Act of 1972 (FEPCA, 1972)]. The 1972 revisions also established the requirement for reregistration of all exist­ ing pesticides within a five-year time frame. When this was not accomplished by 1978, Congress relaxed the timelines [Federal Pesticide Act of 1978 (FPA, 1978)]. Ten years later, following continued delays in completing the rereg­ istration process, Congress once again established specific 1371

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timetables in a five-phase program for all active ingredients registered before November 1, 1984, but this time with tar­ geted funding to support the reregistration activity (FIFRA, 1988). Initially, it was estimated that reregistration would be completed within three to nine years after the enactment of the 1988 amendments. In fact, reregistration activities were completed by the end of 2008. Other important provisions in the FIFRA amendments of 1972 included authorization (1) to require federal regis­ tration of pesticides sold within states, (2) to classify pes­ ticides in general and/or restricted use categories (based upon their inherent acute toxicity) and otherwise regulate their usage (e.g., specify the maximum allowable applica­ tion rate and frequency), and (3) to register establishments that make pesticide products and require them to maintain records, while also being able to inspect these producers, as well as those establishments which hold pesticides for sale, for compliance with the applicable provisions of FIFRA. Amendments to FIFRA in 1975 included the require­ ment for EPA to notify the Secretary of Agriculture in advance of issuing proposals for regulations or to cancel or otherwise change the registration status of a pesticide, and to consider the impact on agriculture when cancellation actions are being considered (FIFRA, 1975). In addition, the seven-member FIFRA Scientific Advisory Panel (SAP) was established to “comment as to the impact on health and the environment” of proposed cancellation actions and regulations. In the years since this provision was added to FIFRA, the SAP has been consulted on many scientific issues reflecting a much broader range of EPA’s pesticide regulatory activities, for example, the proposed classifica­ tion of human cancer potential of a pesticide, the design of a testing protocol, and the risk assessment methodologies developed to address aggregate and cumulative exposure and risk assessment. Congress once again substantially revised FIFRA in 1978 (FPA, 1978). Provisions included granting data sub­ mitters 10 years of exclusive use of their data on new active ingredients while transferring the responsibility for man­ aging the issue of data compensation from the Agency to outside arbitrators. The 1978 revisions also removed most trade secret protection for health and safety data, an early example of public right-to-know. Other changes included the granting of conditional registration authority which allows EPA to approve proposed uses before the full set of supporting data are submitted and reviewed; the estab­ lishment of a procedure for interim administrative review (a process known as Special Review) if “a validated test or other significant evidence raised prudent concerns of unrea­ sonable adverse risk to man or to the environment;” a prohi­ bition against disclosure of data to foreign or multinational pesticide producers; and a requirement for the recognition of the distinction between agricultural and nonagricultural pesticides when processing registration petitions and in setting registration standards and guidelines.

Hayes’ Handbook of Pesticide Toxicology

Among the modifications to FIFRA in 1980 was the provision that the Scientific Advisory Panel (SAP) could create its own subpanels, resolving the limitation in scope of expertise that the smaller, seven-member permanent panel may have on any specific issue (FIFRA, 1980). This allowed expansion of the Panel’s capabilities to pro­ vide more substantive expert scientific peer review of the increasingly diverse and complex issues brought to it by the Agency. The 1980 provisions also required EPA to request SAP comment “as to the impact on health and the environment” of proposed suspension actions (in addition to the requirements for consultation on proposed cancella­ tions) and to set up a process for peer review “with respect to the design, protocols, and conduct of major scientific studies conducted under this Act by the Environmental Protection Agency or by any other Federal agency, any State or political subdivision thereof, or any institution or individual under grant, contract, or cooperative agreement from or with the Environmental Protection Agency” and of “the results of any such scientific studies relied upon by the Administrator with respect to any actions the Administrator may take relating to the change in classification, suspen­ sion, or cancellation of a pesticide.” In 1988, in addition to prescribing the multiphase rereg­ istration program, Congress made other changes to FIFRA, including creation of a “fast-track” registration process for end-use products for “me-too” registrants (“me-too” regis­ trants are those who seek to register products similar to an already registered pesticide product); the necessity of gain­ ing Congressional approval to indemnify registrants holding suspended or cancelled products; and a series of provisions related to the storage, disposal, and transport of suspended or cancelled pesticides and pesticide containers (FIFRA, 1988). Record keeping requirements were expanded to include all registrants and applicants for registration, in addition to those previously required of producers. The Food, Agriculture, Conservation, and Trade Act of 1990 (FACTA, 1990) added requirements for certified pesti­ cide applicators to maintain records of their use of restricteduse chemicals, prohibited registrants of minor-use pesticides from submitting field trial data from geographic areas where the chemical would not be used, and provided discretion to the Administrator to reduce or waive registration fees if the cost would “significantly reduce the availability of the pesticide.” In addition, there were several new requirements related to voluntary cancellation of minor-use pesticides, including a provision for public notice and comment upon a registrant’s application for voluntary cancellation.

62.3  Current state of pesticide regulation in the united states Amendments to FIFRA and FFDCA in 1996 brought both incremental and broad, sweeping changes to the legal

Chapter | 62  Risk Assessment and Risk Management: The Regulatory Process

foundation for pesticide regulation in the United States [The Food Quality Protection Act of 1996 (FQPA, 1996)]. The Food Quality Protection Act (FQPA) represents the outcome of long and complex deliberations to resolve inconsistencies in the existing legislation, both within and between the two statutes.

62.3.1  FIFRA—Key Changes and Additions 62.3.1.1  Emergency Suspension EPA may now suspend a pesticide registration in an emer­ gency situation without simultaneously issuing a notice of intent to cancel, a change from the previous requirement for simultaneous action. A notice of intent to cancel must be issued within 90 days or the suspension will automati­ cally expire. Determination of imminent hazard (to human health or the environment) constitutes grounds for suspen­ sion. This process is invoked to prevent unacceptable risks from occurring during the time required to cancel or oth­ erwise modify the registration of a pesticide. Any action to suspend, cancel or modify the registration of a pesticide under FIFRA must be accompanied by a similar and simul­ taneous action on any associated tolerances under FFDCA.

62.3.1.2  FIFRA Scientific Advisory Panel A Science Review Board to consist of 60 scientists was established to be available to the permanent Panel to assist in the scientific peer reviews conducted by the Panel. Formation of the board complements the earlier modifica­ tion to FIFRA which allowed the Panel to create its own subpanels as needed.

62.3.1.3  Tolerance Reevaluation as Part of Reregistration FIFRA §4 specifies that tolerances and exemptions from tolerances must be reassessed as part of reregistration to determine whether they meet the requirements of the FFDCA. These reassessments must be made as soon as EPA has sufficient information to assess dietary risk, but no later than when it makes product reregistration deci­ sions. These determinations for existing tolerances and exemptions and the need for any additional tolerances or exemptions must be published in the Federal Register and appropriate regulatory action under FIFRA and/or FFDCA is to begin promptly.

62.3.1.4  Registration Renewal The requirement for periodic registration review was intro­ duced. In 2006, EPA initiated its program to reevaluate all pesticides on a regular cycle, with the goal to review each active ingredient every 15 years to make sure that as the

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ability to assess risks to human health and the environment evolves and as policies and practices change, all pesticide products in the marketplace can still be used safely. Formal procedures were established, based upon the Agency’s experience in carrying out the tolerance reassessment pro­ cess mandated by FQPA. Peridiocally, the Agency pub­ lishes, in the Federal Register, the list of pesticides for which the periodic review process will be initiated during the next several fiscal years.

62.3.1.5  Protections for Minor-Use Pesticides, Including Public Health Pesticides A minor use is defined as one in which the pesticide is used on an animal, on a commercial agricultural crop for which the total U.S. acreage is less than 300,000 acres, or for the protection of public health, but does not, on its own, provide sufficient economic incentive to support the costs of registration or reregistration. In addition, a minor use pesticide must play a significant role in managing pest resistance or in an integrated pest management (IPM) pro­ gram. There also must be a lack of efficacious alternatives or the alternatives must pose a greater risk to human health or the environment than does the pesticide under evalua­ tion. Many minor-use crops are fruits and vegetables, which are significant components of the human diet. The law provides additional incentives for the develop­ ment and maintenance of minor use registrations in a num­ ber of ways: extension of time to generate residue data and for exclusive use of these data; greater flexibility to waive data requirements; the option to waive some or all of the fees usually charged to support and maintain registration; and expedited review of minor-use applications by the Agency. None of these provisions would apply, however, if the minor use is determined to pose unreasonable risks or if the lack of data would significantly delay EPA decisions. FQPA establishes a USDA revolving grant program and a program for support of public health pesticides to be implemented jointly by the Public Health Service of HHS and EPA. By virtue of instituting this program, the federal government bears the cost of developing the required data to support the registration and reregistration of the pub­ lic health use, as it does for minor-use pesticides used on agricultural crops under the USDA Inter-Regional Project Number 4 (IR-4) program. A public health pesticide is defined in FIFRA §2(nn) as “any minor use pesticide prod­ uct registered and used predominantly in public health programs for vector control or for other recognized health protection uses, including the prevention of viruses, bacte­ ria, or other microorganisms (other than viruses, bacteria, or other microorganisms on or in living man or other animal) that pose a threat to public health.” A “vector” is defined in FIFRA §2(oo) as “any organism capable of transmit­ ting human discomfort or injury, including mosquitoes, flies, fleas, cockroaches, or other insects and ticks, mites,

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or rats.” Perhaps the most notable example of a public health pesticide is DDT, still used in some parts of the world, but not the United States, for the control of mosquitoes bearing the malaria vector.

62.3.1.6  Antimicrobial Pesticides FIFRA §2(mm) defines an antimicrobial pesticide as one which is intended to “disinfect, sanitize, reduce, or miti­ gate growth or development of microbiological organisms” or “protect inanimate objects, industrial processes or sys­ tems, surfaces, water, or other chemical substances from contamination, fouling, or deterioration caused by bacteria, viruses, fungi, protozoa, algae, or slime” and, in this use, is exempt from a tolerance under FFDCA §408. Wood preser­ vatives, antifouling paints, agricultural fungicides, aquatic herbicides, and liquid chemical sterilants intended for use on critical or semicritical medical devices as defined under FFDCA are not included within the definition. FQPA contains special provisions for antimicrobial pesticides, essentially removing them from the shadow of pesticides intended for agricultural and other nonfood uses and prompting more focused attention on facilitating more timely registration decisions. The law requires the identi­ fication, evaluation, and implementation of reforms to the registration process for this class of pesticides to reduce review periods, providing explicit goals in number of days depending upon the action requested (i.e., a new use of an already-registered active ingredient; a new product; “me-too’s”; and amendments to existing uses). A separate administrative unit has been established within the Office of Pesticide Programs that deals only with the registration, reregistration, and Special Review processes for antimicro­ bial pesticides. A separate section in 40 CFR 158 describes the data requirements necessary to support registration or reregistration for these products.

62.3.1.7  Reduced Risk or “Safer” Pesticides In the early 1990s, the Office of Pesticide Programs set up a system by which reduced risk or “safer” pesticides would be given priority attention in the registration pro­ cess. Guidelines governing the expedited review of con­ ventional and biological pesticides were issued in 1997 [U.S. Environmental Protection Agency (U.S. EPA, 1997a)]. FQPA provided the statutory mandate for continu­ ing this expedited consideration of applications for pesti­ cides which meet one or more of the criteria for a reduced risk pesticide. A pesticide qualifies for expedited review as a reduced risk pesticide if its use “may reasonably be expected to accomplish one or more of the following: (1) reduce the risks of pesticides to human health; (2) reduce the risks of pesticides to nontarget organisms; (3) reduce the potential for contamination of groundwater, surface water, or other valued environmental resources; and (4) broaden

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the adoption of IPM strategies, or make them more available or effective “ [FIFRA §3(c)(10)(B)].

62.3.1.8  Data Collection The keystone of FQPA is the inclusion of special provisions for infants and children. Title III of FQPA addresses data collection activities to assure the health of infants and chil­ dren. It states that USDA, in cooperation with FDA and/or EPA, “shall coordinate the development and implementa­ tion of survey procedures to ensure that adequate data on food consumption patterns of infants and children are col­ lected”; shall ensure there will be improved data collection on the occurrence of pesticide residues in foods, particularly those most likely to be consumed by infants and children; and shall evaluate the current status of pesticide usage infor­ mation and move to improve usage information gathering activities. Information in all three of these areas is critical to the conduct of credible and accurate estimates of risk from exposure to pesticide residues in the diet.

62.3.2  FFDCA—Key Changes and Additions The most significant changes in pesticide regulation result­ ing from the passage of FQPA impact the tolerance-setting process described in FFDCA. Definitional and process changes were mandated and the factors that are to be con­ sidered when conducting risk assessments and making risk management decisions were expanded.

62.3.2.1  The Delaney Clause Until FQPA was passed, pesticide residues in processed foods were considered to be “food additives” regulated under FFDCA §409. If a pesticide residue was expected to exceed the level which was allowed under a §408 toler­ ance for the raw agricultural commodity, it became neces­ sary to establish a separate food additive regulation for the processed food under §409. However, the Delaney clause in §409 prohibits the establishment of food additive regula­ tions for any substance “if it is found to induce cancer when ingested by man or animal, or if it is found, after tests which are appropriate for the evaluation of the safety of food addi­ tives, to induce cancer in man or animal. . . .” Under FQPA, pesticide residues are excluded from the definition of “food additive.” Thus, the Delaney clause is no longer applicable to pesticide residues in processed foods. All pesticide resi­ dues, whether in raw or processed foods, are regulated only under FFDCA §408, which does not contain the prohibition against setting tolerances for carcinogens.

62.3.2.2  Definition of “Safe” Under FFDCA §408(b)(2)(A), the standard for establish­ ing a tolerance is based on whether the tolerance is “safe.”

Chapter | 62  Risk Assessment and Risk Management: The Regulatory Process

To be “safe” means that there is “a reasonable certainty that no harm will result from aggregate exposure to the pesticide chemical residue, including all anticipated dietary exposures and all other exposures for which there is reliable information.” This definition is consistent with the stan­ dard applied historically to nonpesticide food additives and color additives by the Food and Drug Administration. For threshold effects (i.e., those effects for which a level can be identified that would not be expected to cause or contrib­ ute to any adverse human health consequences), the safety standard is satisfied if the aggregate exposure is lower than the no-effect level by “an ample margin of safety.” Traditional regulatory policy states that, as a default, exposure estimated to be at or below a level 100-fold lower than the critical no-effect level identified in the ani­ mal toxicology database would meet the safety standard. For non-threshold effects (i.e., those for which no no-effect level can be identified), a pesticide will satisfy the safety standard if the increased lifetime risk, expressed as a prob­ ability, is “negligible.” Traditionally, “negligible” has been defined as being no greater than a one-in-a-million excess lifetime risk for nonoccupational exposures.

62.3.2.3  Aggregate Exposure Aggregate exposure is defined as that which occurs from all food uses for the pesticide, as well as from exposure that occurs from all nonoccupational sources. This would include exposures from drinking water, nonfood pesti­ cidal uses (e.g., lawn and garden use or indoor residential, school, or public building applications) and those expo­ sures that may result from nonpesticidal uses (e.g., as a human pharmaceutical or a hazardous waste site contami­ nant). Guidance for conducting aggregate exposure and risk assessment has been developed by EPA (U.S. EPA, 2001). Among the factors that must be taken into account when establishing, modifying, leaving in effect, or revok­ ing a tolerance or an exemption from a tolerance is the risk that may ensue from the aggregate exposure to the pesticide under evaluation. A tolerance represents a single pesticide– use combination. That is, one tolerance would be needed if Chemical X were to be used on potatoes. A separate toler­ ance would be required if Chemical X also were to be used on lettuce. Therefore, when making a decision with regard to any one use, EPA must consider the exposure and risk that would occur not only as a consequence of that particu­ lar use, but also all other existing and proposed food and nonoccupational, nonfood uses.

62.3.2.4  Common Mechanism of Toxicity and Cumulative Risk Assessment Another factor that must be taken into account when estab­ lishing, modifying, leaving in effect, or revoking a toler­ ance or an exemption from a tolerance is the cumulative

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effects of the pesticide under evaluation and other sub­ stances with which it may share a common mechanism of toxicity. “Other substances” are not just other pesticides, but may be drugs, commodity chemicals, or environmen­ tal contaminants. When a common mechanism finding is made, a cumulative risk assessment is to be considered. The first step is to determine the need for a cumulative risk assessment. This is done by conducting a hazard assess­ ment for that pesticide. In the course of performing this hazard assessment, information would come to light to suggest that this pesticide may share a common mecha­ nism with at least one other pesticide. This process con­ tinues until all likely pesticide candidates are identified. After it is concluded that two or more pesticides are can­ didates for a common mechanism group, an effort is made to determine if reliable information exists to suggest that any nonpesticides also may share the same mechanism of toxicity. EPA has developed criteria by which to judge whether substances share a common mechanism of toxic­ ity (U.S. EPA, 1999). At the conclusion of the process to determine those substances which actually do share a com­ mon mechanism of toxicity, those remaining substances are subjected to cumulative risk assessment. It is possible that the final cumulative risk assessment may not include all of the substances that constituted the original common mech­ anism group. Modifications to group membership would be informed by the results of the individual aggregate exposure assessments that also would be conducted on each original candidate for the common mechanism group. The nature, magnitude, and timing of exposure to each substance in the aggregate and how the exposures (or their biological con­ sequences) to individual substances overlap become criti­ cal factors in determining the final group to be included in the cumulative risk assessment. EPA has developed guid­ ance for the conduct of cumulative risk assessments (U.S. EPA, 2002a). To date, four groups of substances have been subjected to a cumulative risk assessment: 31 cholinesteraseinhibiting organophosphorus insecticides, 10 cholinesteraseinhibiting N-methyl carbamate insecticides, three triazine herbicides (atrazine, propazine, simazine) and their three chlorinated degradates and two chloroacetanilide herbicides (alachlor and acetochlor). At the present time, research is underway to determine whether or not any of the mem­ bers of the pyrethroid class of insecticides share a common mechanism of toxicity and, thus, would be candidates for a cumulative risk assessment.

62.3.2.5  Special Considerations for Infants and Children When making tolerance decisions, EPA also must imple­ ment several new requirements related to assuring the safety of infants and children. The Agency must assess (aggregate) risk based upon available information about: (1) dietary consumption patterns that are likely to yield

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disproportionately higher exposures or risks; (2) special susceptibilities to pesticides, including neurological dif­ ferences between infants and children and adults, and the effects of in utero exposure to pesticide chemicals; and (3) the cumulative effects of the pesticide residues and other substances that have a common mechanism of toxic­ ity. The “reasonable certainty of no harm” safety standard must be ensured and a specific safety determination for infants and children must be made. The provision that has prompted the most controversy and has had the greatest impact upon the risk assessment and regulatory decision-making process under FQPA is the obligatory application of an additional safety factor. FFDCA §408(b)(2)(C) states that “in the case of thresh­ old effects . . . an additional tenfold margin of safety for the pesticide chemical residue and other sources of expo­ sure shall be applied for infants and children to take into account potential pre- and postnatal toxicity and complete­ ness of data with respect to exposure and toxicity to infants and children.” A different margin of safety may be used only if, on the basis of reliable data, such a margin will be safe. It should noted that any different margin of safety could be greater or lesser than the default 10X. Since the passage of FQPA, EPA has developed a series of policy guidance documents, articulating its approach to implementing the “FQPA Safety Factor” provision of the law. The most current thinking on this topic can be found in the document entitled Determination of the Appropriate FQPA Safety Factor(s) in Tolerance Assessment (U.S. EPA, 2002b). This document describes when FQPA safety factor decisions are needed; what the FQPA 10X safety factor is “in addition to”; how to judge the completeness of the toxicology and exposure databases; when a data­ base uncertainty factor greater than 1X is applied; how to determine, and account for, the degree of concern for pre- and postnatal toxicity; and the process for determina­ tion of the appropriate FQPA safety factor(s). Additional draft guidance on the application of the FQPA safety fac­ tor in cumulative risk assessments also is available (U.S. EPA, 2002c).

62.3.2.6  Consumer Right-to-Know FFDCA §408(o) states that EPA shall publish, and provide to large grocery stores, a publication which describes the risks and benefits of pesticide residues on food purchased in those stores by consumers, a listing of those pesticides for which the limited benefits-based tolerances have been issued, and recommendations for consumers on how they can reduce dietary exposures to pesticides in a manner con­ sistent with maintaining a healthy diet. EPA developed and distributed a brochure, Pesticides and Food: What You and Your Family Need to Know, to more than 30,000 grocery stores. Copies also went to public health officials, libraries, and the medical community. Over six million copies are

Hayes’ Handbook of Pesticide Toxicology

in circulation. A discussion of the topics addressed in the brochure also can be found on EPA’s Office of Pesticide Programs’ Web site (www.epa.gov/pesticides/food).

62.3.2.7  Estrogenic Substances Screening Program FFDCA §408(p) states the EPA shall “develop a screen­ ing program, using appropriate validated test systems and other scientifically relevant information, to deter­ mine whether certain substances may have an effect in humans that is similar to an effect produced by a natu­ rally occurring estrogen, or such other endocrine effect as the Administrator may designate.” The Agency was given two years to develop the screening program, another year to implement it, and four years to report on its find­ ings. The Agency established an advisory committee to assist it with the development of the program. In August 1998, the committee released its recommendations, most of which were adopted by the Agency. Implementation of the Endocrine Disruptor Screening Program (EDSP) began soon thereafter, with the initial focus being on the standardization and validation of the proposed components of the screening and testing batteries that would make up the screening program. As of late 2009, EPA had: 1) devel­ oped, validated and selected the components of the Tier 1 screening battery, soliciting review and comment from the FIFRA Scientific Advisory Panel on relevant techni­ cal issues; 2) published its approach for selecting the first 50–100 chemicals to be screened; 3) issued a draft list of 73 chemicals for the first round of screening and testing; 4) published implementation policies and procedures by which testing orders will be issued, and 5) issued test orders for the first 67 chemicals. Additional details on the program and the Report to Congress (August 2000) can be found on the Web site of EPA’s Office of Science Coordination and Policy (www.epa.gov/endo).

62.3.2.8  Tolerance Reassessment FQPA required EPA to reevaluate all 9721 tolerances and tolerance exemptions in effect on the day before enactment of the act on August 4, 1996. A schedule was imposed that was to assure that 33% of such tolerances and exemptions were reviewed within three years of enactment; a second 33% within six years; and the remaining number within 10 years. Priority was to be given to those tolerances or exemptions that appeared to pose the greatest risk to pub­ lic health (i.e., review the “worst first”). EPA divided all chemicals into three groups, with Priority Group 1 con­ taining the pesticides that appeared to pose the greatest risks. Priority Group 1 was made up of several subgroups: organophosphorus compounds (OPs); carbamates; pes­ ticides which had previously been characterized as prob­ able human carcinogens (Groups B1 and B2) according

Chapter | 62  Risk Assessment and Risk Management: The Regulatory Process

to EPA’s classification scheme published in its 1986 cancer risk assessment guidelines; high-hazard food-use inert ingredients; and any chemicals that exceeded their ref­ erence dose (RfD) by unacceptable levels. Priority Group 2 contained those pesticides characterized as possible human carcinogens (Group C, in accordance with the 1986 can­ cer classification scheme) and all reregistration chemicals which remained unfinished in 1996 (i.e., of those registered before 1984). Priority Group 3 contained all remaining preFQPA chemicals for which reregistration eligibility deci­ sions already had been made by August 1996; all remaining post-1984 chemicals; biopesticides; and the rest of the food-use inert ingredients. EPA published a notice in the Federal Register on August 4, 1997, outlining its plans for carrying out the tolerance reassessment process and identi­ fying the individual substances in each of the three priority groups (U.S. EPA, 1997a). EPA fully met the 33% target by August 3, 1999 and the 66% target by August 3, 2002. At the time of the final deadline (August 3, 2006), reassess­ ments of greater than 99% of the 9721 tolerances or exemp­ tions from a tolerance had been completed. Reassessment of the remaining 84 tolerances (all associated with uses of the N-methyl carbamate insecticides which also were being evaluated in a cumulative risk assessment based upon common mechanism of toxicity) was completed in September, 2007.

62.4  Current regulatory process The registration of pesticides in the United States is bounded by a structure defined by congressional legislation, as inter­ preted in formal regulations and other less-formal articu­ lations of policy and practice. Proposals for changes and final changes to the regulations are published in the Federal Register. All final, formal regulations also can be found in the Code of Federal Regulations (40 CFR Parts 150–189). The CFR is updated annually to reflect any changes in the regulations that may have been finalized during the year. Daily issues of the Federal Register and the current CFR can be found on the Web site of the U.S. Government Printing Office (http://www.gpo.gov/). Frequently, statements of policy related to pesticide regulation are published by EPA’s Office of Pesticide Programs (OPP) as Pesticide Registration (PR) Notices. These PR Notices as well as other documents articulating OPP’s regulatory and risk assessment policies and prac­ tices can be found on OPP’s Web site (http://www.epa.gov/ pesticides/). Because regulatory approaches and practices are continually updated as the state-of-the-science and its interpretation evolve, prospective pesticide registrants are strongly encouraged to meet with pesticide officials before proceeding with data generation and submission of peti­ tions for registration or before executing changes in the registration status of their products.

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62.4.1  Registration Registering a pesticide product for use in the United States under the regulations of the Federal Insecticide Fungicide and Rodenticide Act is equivalent to acquiring a federal license to sell or distribute a product in commerce. To do so without EPA approval is a federal crime. In theory, all pesticide products destined for use in the United States must be registered. A pesticide prod­ uct is generally made up of more than one constituent. It may include one or more “active” ingredient(s) and one or more “inert” ingredient(s). A pesticide active ingredient is defined as “(1) any substance or mixture of substances intended for preventing, destroying, repelling, or mitigating any pest; (2) any substance or mixture of substances intended for use as a plant regulator, defoliant, or dessi­ cant, and (3) any nitrogen stabilizer” [FIFRA §2(u)]. An “inert ingredient” is any substance or group of similar sub­ stances, other than the active ingredient, which is inten­ tionally included in a pesticide product. Both the active ingredient(s) and the formulation(s) which constitute the product(s) are subject to registration requirements. If the intended use of the product includes application to agricultural commodities destined for human or animal consumption, a tolerance or exemption from a tolerance also must be granted under FFDCA §408. New animal drugs, animal feeds containing a new ani­ mal drug, and liquid chemical sterilants for use on criti­ cal or semi-critical medical devices are excluded from the definition of “pesticide” and are regulated by the Food and Drug Administration. Although FIFRA §3 requires the registration of all materials which meet the definition of “pesticide” (i.e., either an active or inert ingredient), this section along with several other sections of the law pro­ vides for exemptions. For instance, exemptions may be granted if the material is being transferred from one site of an establishment to another when both are operated by the same producer, if it is regulated by another federal agency (e.g., certain biological control agents and human drugs), if it is “of a character which is unnecessary to be subject to this Act” [FIFRA §25(b)], or if it is in the preregistration status of having been granted an experimental-use permit under FIFRA §5 or an emergency use under FIFRA § 18. Approval of a registration is dependent upon the suc­ cessful fulfillment of a series of data requirements, among other factors. The number and types of studies to be con­ ducted vary with the intrinsic chemistry, anticipated inher­ ent toxicity, and proposed use pattern of the pesticide. Pesticides of conventional chemistry proposed for use on agricultural commodities generally require the greatest amount of information, whereas nonfood-use conventional chemicals, antimicrobials, and biopesticides such as micro­ bials and biochemicals generally require less. Part 158 of 40 CFR presents the regulatory roadmap specifying the types and amounts of data and other information needed

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by EPA to decide whether or not to approve an application for a new or amended registration or reregistration under FIFRA §3, for an experimental-use permit under FIFRA §5, or for a emergency exemption under FIFRA § 18. The data requirements specified in this part cover the areas of prod­ uct chemistry, toxicology for human health and terrestrial mammals, wildlife and aquatic toxicology, nontarget insects (e.g., honey bees), environmental fate, aerial drift evaluation, reentry protection (primarily in the occupational setting), plant protection, product performance, residue chemistry (for food uses), and biochemical and microbial pesticides. Some of these kinds of data are always required for the evaluation of some or all types of products. Other kinds of data are required only under certain conditions, that is, if the product’s proposed pattern of use, the results of earlier stud­ ies, or other circumstances warrant the development of such data. For example, the acute delayed neurotoxicity study in the hen is required only if the pesticide is an organophos­ phate or a metabolite thereof and causes inhibition of ace­ tylcholinesterase or is structurally related to a substance that is known to cause delayed neurotoxicity. Another example would be the requirement to develop residue chemistry data only if the pesticide is proposed for use on food crops. Some, but not all, subparts of Part 158 have been modi­ fied since their original promulgation in 1984. Other data requirements have been added without the benefit of formal promulgation of regulations. This has occurred because the state-of-the-science has evolved and matured in many areas in the intervening years. The Office of Pesticide Programs has modified many of its data requirements, communicat­ ing these changes via Pesticide Regulation Notices and other written materials. Thus, although the current Part 158 establishes data requirements that are applicable to various general use patterns, some unique aspect of a proposed use and/or the possibility of modification of the original data requirements for any use, general or unique, argues strongly for consultation between the prospective regis­ trant and the Agency before beginning any data generation or information development.

62.4.2  Reregistration and Registration Renewal Four times, over approximately 25 years (1972, 1978, 1988, 1996), Congress acted to require EPA to update the registration status of existing pesticide products. The most prescriptive directive was introduced in the amendments to FIFRA in 1988. Implementation of the reregistration scheme articulated at that time was completed by the end of 2008, having been made more complex by the require­ ment to reassess all of the tolerances and tolerance exemp­ tions in place when FIFRA and FFDCA were amended in the Food Quality Protection Act of 1996. The reregistration directives in the 1972 and 1978 amendments to FIFRA covered those pesticides registered

Hayes’ Handbook of Pesticide Toxicology

up to that point. The accelerated reregistration program under FIFRA 1988 (FIFRA §4) encompassed all pesticide active ingredients initially registered before November 1, 1984. At that time, there were approximately 1150 active ingredients and over 20,000 product formulations regis­ tered in the United States. Because many of these 1150 active ingredients were related to one other (e.g., different salts of the same substance, such as sodium and calcium hypochlorite), they were organized into about 600 “cases” or groups of related pesticide active ingredients. These 600 cases were divided into four lists: List A—List A, which contains most of the food use pesticides, consists of the 194 chemical cases (or 350 indi­ vidual active ingredients) for which EPA had issued reg­ istration standards prior to FIFRA 1988. Each registration standard document summarized the data available for a pesticide, called in any additional studies needed for rereg­ istration, and required necessary product labeling changes. Lists B, C, and D—The remaining pesticides requiring reregistration, and for which no registration standard had been developed in previous reregistration attempts, were divided into three lists based on their potential for human exposure and other factors, with List B containing pesti­ cides of greater concern, List C pesticides of lesser concern and List D pesticides of least concern. Some of the classi­ fication criteria included the potential or known occurrence of residues in food or drinking water, significance of out­ standing data requirements, potential for worker exposure, Special Review or restricted-use status, and unintended adverse effects on animals and plants. FIFRA 1988 established a reregistration process con­ sisting of five phases, with time frames and responsibilities for both EPA and the pesticide producers or registrants. The pesticides on Lists B, C, and D went through all five phases. Because EPA had already substantially reviewed them under the Registration Standards program, the List A pesticides moved directly to Phase 5. Phase 1, list active ingredients—As required, EPA published Lists A, B, C, and D within 10 months of the passage of FIFRA 1988 (by October 24, 1989) and asked registrants of these pesticides whether they intended to seek reregistration. Phase 2, declare intent and identify studies—Phase 2 required registrants to notify EPA whether or not they intended to reregister their products; to identify and com­ mit to providing necessary new studies; and to pay the first installment of the reregistration fee. During this phase, EPA issued guidance to registrants for preparing their Phase 2 and Phase 3 responses. Phase 2 activities were completed in 1990. Nearly 250 cases, which included nearly half of the existing products, did not proceed past this phase, as the registrants chose not to support their con­ tinued registrations. Phase 3, summarize studies—During Phase 3, follow­ ing EPA guidance, registrants were required to submit summaries and reformat acceptable studies, “flag” studies

Chapter | 62  Risk Assessment and Risk Management: The Regulatory Process

indicating adverse effects, recommit to satisfying all appli­ cable data requirements, and pay the final installment of the reregistration fee. Phase 3 ended in October 1990. Phase 4, EPA review and data call-in—During Phase 4, EPA reviewed all Phase 2 and Phase 3 submissions and required registrants to meet any unfulfilled data require­ ments within four years. Phase 4 was completed in 1993. Phase 5, reregistration decisions—In this final phase, which was completed by the end of 2008, EPA reviews all the studies that have been submitted and decides whether or not the active ingredient(s) and the pesticide products containing the active ingredient(s) are eligible for reregis­ tration—whether the data base is substantially complete, and whether or not the pesticide causes unreasonable adverse effects to humans or the environment when used according to product labeling. EPA also considers whether the pesticide meets the new safety standard of the FQPA and conducts tolerance reassessment for those pesticides which have food uses. The results of the Agency’s review are presented in a Reregistration Eligibility Decision (RED) document. Products containing the pesticide active ingredient are reregistered after certain product-specific data and revised labeling are submitted and approved. All of the active ingredients in a pesticide product must be eli­ gible before the product is considered to be reregistered. Because of the FQPA requirement that pesticides sharing a common mechanism of toxicity with other substances (with or without pesticidal uses of their own) shall be eval­ uated for inclusion in a cumulative risk assessment with these other substances, the Agency issued Interim REDs (IREDs) for some pesticides, reflecting its judgment con­ cerning reregistration eligibility based solely on the indi­ vidual pesticide’s aggregate risk assessment, but reserving judgment on full reregistration eligibility until the cumula­ tive risk assessment process was completed. The timetable for reregistration that had been estab­ lished in response to the congressional directives in FIFRA 1988 required modification following the passage of FQPA. Among the new dimensions added in FQPA was the require­ ment to reassess, within a 10-year time frame, all previously granted tolerances and exemptions from the requirement for a tolerance against the new safety standard that FQPA had established. This directive applied to tolerances (and exemp­ tions) for all food-use pesticides, without regard to the date of their original registration. The consequences of this direc­ tive were, among others, that those food-use pesticides for which REDs already had been completed under the FIFRA 1988 reregistration process needed to be revisited for tol­ erance reassessment. Some of these also would need to be considered for inclusion in a cumulative risk assessment. Because EPA used the reregistration program to accomplish tolerance reassessment, the timetable for completion of reregistration was extended to encompass the 10-year time frame for tolerance reassessment, 1996–2006. By the end of fiscal year 2008 (September 30), reregis­ tration of over 98% of the original 613 cases was completed.

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The remainder were completed by the end of the calendar year. Persons interested in obtaining details on the status of individual cases are referred to EPA’s Office of Pesticide Programs Web site, where status reports and other materi­ als on reregistration can be found at (http://www.epa.gov/ pesticides/reregistration/index.htm). FQPA also required EPA to establish a new registration review program. This new program obligates EPA to review every registered pesticide on a 15-year cycle. This new pro­ gram includes all pesticides registered since November 1, 1984, as well as those that have been through earlier rereg­ istration processes. Implementation of such a program assures that pesticides are being reviewed periodically and updated to meet current scientific and regulatory standards. EPA initiated its Registration Review program in 2006 while the original reregistration program was still ongoing. The Agency has issued a summary of the planned schedule for the initiation of activities on a series of conventional, antimicrobial, biochemical and microbial pesticides begin­ ning in FY 2009 through FY 2012 (http://www.wpa.gov/ oppsrrd1/registration_review/schedule.htm). Other details on the Registration Review program can be found on EPA’s Office of Pesticide Programs’ website (http://www.wpa .gov/oppsrrd1/registration_review/index.htm).

62.4.3  Special Review EPA not only has the authority to register pesticides, but also to cancel, suspend, or modify the registration of any pesticide or use of such pesticide that the Agency has determined to have the potential to “cause unreason­ able adverse effects on the environment” [FIFRA §6(b)]. FIFRA §(2)(bb) states that, in making a final judgment whether to cancel or modify the conditions of registration of a pesticide or any of its uses, the Agency must weigh the potential for adverse effects against the costs and ben­ efits (“economic, social, and environmental”) derived from the use(s) of the pesticide. Dietary risks must be judged against the “reasonable certainly of no harm” safety stan­ dard under Section 408 of FFDCA. Risks related to use of public health pesticides are weighed against the health risks such as those from the diseases transmitted by the vector to be controlled by the pesticide. The formal procedure for conducting the necessary reg­ ulatory assessment under FIFRA §6 is commonly known as Special Review. Regulations governing the Special Review process are articulated in 40 CFR Part 154. The formal Special Review process which includes a FIFRA §6 cancellation or suspension hearing is resource-intensive and time-consuming. In practice, the Agency has more often used less formal procedures to achieve the same goal of reducing the potential risks to acceptable limits. In more recent times, the reregistration process has been the princi­ pal mechanism for negotiating changes in the registration status of pesticides. It is more efficient and, whereas the

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formal Special Review process generally is a dialogue only between the Agency and the registrant(s) (and, sometimes, the hearing judge), with user groups free to submit benefits and other economic information to USDA, the less formal, and less time-consuming, procedure encourages and sup­ ports a more active role for user groups and other inter­ ested stakeholders, such as public interest groups and the public health community. Persons interested in obtaining details on the status of chemicals in Special Review are referred to EPA’s Office of Pesticide Programs Web site, where its latest report (March 2000) is available (http://www.epa.gov/oppsrrd1/special_ review/sr00status.pdf). Risk reduction measures taken as a result of assessments conducted during the reregistra­ tion process are detailed in individual REDs and IREDs and summarized periodically in status reports. These can be found on EPA’s Office of Pesticide Programs Web site, where status reports and other materials on reregistration can be found (http://www.epa.gov/pesticides/reregistra­ tion/index.htm) and/or Regulations.gov, the public face of the Federal E-Government eRulemaking Program.

62.5  Web Sites Government Printing Office Code of Federal Regulations and Federal Register (http:// www.gpo.gov/) Federal E-Government eRulemaking Program (http://www. regulations.gov) U.S. Environmental Protection Agency Consumer Right-to-Know Pesticides and Food: What You and Your Family Need to Know (http://www.epa.gov/pesticides/food) Endocrine Disruptor Screening Program (http://www.epa. gov/scipoly/oscpendo/index.htm) General information on the pesticide regulatory program (http://www.epa.gov/pesticides/) Registration Review (http://www.wpa.gov/oppsrrd1/regis­ tration_review/index.htm). Reregistration (http://www.epa.gov/pesticides/reregistra­ tion.htm) Special Review (http://www.epa.gov/docs/SpecialReview/ sr00status.pdf)

References Code of Federal Regulations. (1990). 40 CFR Parts 150–189. Protection of the environment. Subchapter E—Pesticide Programs.

Hayes’ Handbook of Pesticide Toxicology

FACTA. (1990). Food, Agriculture. Conservation, and Trade Act of 1990, Public Law No. 101-624, sees. 1491–1496. 104 Stat. 3359. FEPCA. (1972). Federal Environmental Pesticide Control Act of 1972, Public Law 92-516, 86 Stat. 973. FFDCA. (1954). Federal Food, Drug, and Cosmetic Act of 1954 Miller Amendment to FFDCA §408, Public Law No. 518, 68 Stat. 511. FFDCA. (1958). Federal Food, Drug, and Cosmetic Act of 1958 Food Addi­ tives Amendment to FFDCA §409, Public Law 85-929, 72 Stat. 1785. FIA. (1910). Federal Insecticide Act of 1910, Chap. 191, 36 Stat. 331. FIFRA. (1947). Federal Insecticide, Fungicide, and Rodenticide Act of 1947, Public Law No. 80-104, 61 Stat. 163. FIFRA. (1964). Federal Insecticide, Fungicide, and Rodenticide Act of 1964, Public Law 88-305, 78 Stat. 190. FIFRA. (1975). Federal Insecticide, Fungicide, and Rodenticide Act of 1975, Public Law No. 94-140, 89 Stat. 751. FIFRA. (1980). Federal Insecticide, Fungicide, and Rodenticide Act of 1980, Public Law No. 96-539, 94 Stat. 3194. FIFRA. (1988). Federal Insecticide, Fungicide, and Rodenticide Act Amendments of 1988, Public Law No. 100-532, 102 Stat 2654. FPA. (1978). Federal Pesticide Act of 1978, Public Law No. 95-396, 92 Stat. 819. FQPA. (1996). Food Quality Protection Act of 1996, amending the Federal Insecticide, Fungicide, and Rodenticide Act and the Federal Food, Drug, and Cosmetic Act, Public Law No. 104-170, 11 Stat 1513. Nixon, R. M., President. (1970). Reorganization Plan No. 3 of 1970. 40 CFR pt 1 and Fed. Reg. 35, 15623. U.S. Environmental Protection Agency. (1997a). Guidelines for Expedited Review of Conventional Pesticides under the Reduced-Risk Initiative and for Biological Pesticides. Pesticide Registration Notice 97-3, September 4, 1997. (http://www.epa.gov/PR_Notices/pr97-3.html) U.S. Environmental Protection Agency (1997b). Raw and processed food schedule for pesticide tolerance reassessment. Fed. Reg. 62(149), 42019–42030 August 4, 1997. U.S. Environmental Protection Agency. (1999). Guidance for Identifying Pesticide Chemicals and Other Substances Which Have a Common Mechanism of Toxicity. Office of Pesticide Programs. February 1999. (http://www.epa.gov/fedrgstr/EPA-PEST1999/February/Day-05/ 6055.pdf) U.S. Environmental Protection Agency. (2001). General Principles for Performing Aggregate Exposure and Risk Assessments. Office of Pesticide Programs. November, 2001 (http://www.epa.gov/pesticides/ trac/science/aggregate.pdf) U.S. Environmental Protection Agency. (2002a). Guidance on Cumulative Risk Assessment of Pesticide Chemicals That Have a Common Mechanism of Toxicity. Office of Pesticide Programs. January, 2002. (http://www.epa.gov/pesticides/trac/science/cumulative_guidance.pdf). U.S. Environmental Protection Agency. (2002b). Determination of the Appropriate FQPA Safety Factor(s) in Tolerance Assessment. Office of Pesticide Programs. February, 2002. (http://www.epa. gov/oppfead1/trac/science/determ/pdf). U.S. Environmental Protection Agency. (2002c). Consideration of the FQPA Safety Factor and Other Uncertainty Factors in Cumulative Risk Assessment of Chemicals Sharing a Common Mechanism of Action. Office of Pesticide Programs. February 2002 draft. (http:// www.epa.gov/oppfead1/trac/science/consid_draft.pdf)

Chapter 63

Perceptions of Pesticides as Risks to Human Health Paul Slovic Decision Research, Inc.

63.1  Introduction Public perceptions of risk have been studied systematically for more than 20 years, within the United States and in other countries. Throughout that time period, the use of pesticides has been perceived as one of the most risky activities pursued by human societies. It is difficult to pinpoint the origins of these perceptions but certainly Rachel Carson’s book, Silent Spring, first published in 1962, has played an important role. In Carson’s story, pesticides are singled out as among the most potent “elixirs of death.” Referring to synthetic pesticides, Carson observed, “They have immense power not merely to poison but to enter the most vital processes of the body and change them in sinister and often deadly ways” (Carson, 1962, p. 25). Ironically, as will be shown, the heavy reliance on testing of chemicals with animals and the quantitative risk assessment that has developed during this same time period may have reinforced and maintained the public’s fears of pesticides and other chemicals.

63.2  Risk-perception studies One of the first quantitative studies of risk perception took place in the United States in the late 1970s and early 1980s. This research showed that perceptions of risk can be described in terms of numerous characteristics or dimensions. Figure 63.1, for example, presents a spatial display of hazards within a perceptual space derived from more than 40,000 individual judgments. The factors in this space reflect the degree to which a hazard is perceived to be known or understood (vertical dimension) and the degree Hayes’ Handbook of Pesticide Toxicology Copyright © 2001 Elsevier Inc. All rights reserved

to which it evokes perceptions of dread, uncontrollability, and catastrophe (horizontal dimension). Research has also demonstrated that social response to risk is closely related to the position of a hazard within this space. The further to the right a hazard appears, the higher its perceived risk, the more people want to see its current risks reduced, and the more they want to see strict regulation employed to achieve reduced risk. Media coverage appears to be most extensive and intense when something goes wrong in the upper right-hand quadrant of the space, in an activity whose risks are seen as being poorly understood, evoking dread, and potentially leading to a catastrophe. In this light, it is interesting to see that pesticides fall in the problematic upper right quadrant of the space, reflecting the fact that respondents in this study characterized them as poorly known or understood, delayed in effect, relatively new, uncontrollable, evoking dread, catastrophic, fatal, inequitable, and posing high risk to future generations. Their location in this space is not too distant from activities related to the use of nuclear power. Whereas public judgments of risk seem closely related to the characteristics that define the space in Fig. 63.1, expert’s judgments of risk are not closely related to any of these various risk characteristics. Instead, experts appear to see riskiness as synonymous with expected annual mortality. As a result, many conflicts in society over “risk” may result from experts and laypeople having different definitions of the concept. In this light, it is not surprising that expert recitations of “risk statistics” often do little to change people’s attitudes and perceptions. In addition to constructing spatial displays such as that in Fig. 63.1, research has compared perceptions of risk and 1381

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Figure 63.1  Location of 81 hazards in a two-factor space derived from the relationships among 15 risk characteristics. Each factor is made up of a combination of characteristics, as indicated by the lower diagram. [Slovic, P., Fischhoff, B., and Lichtenstein, S. (1985). Characterizing perceived risk. In “Perilous Progress: Technology as Hazard” (R. W. Kates, C. Hohenemser, and J. X. Kasperson, eds.), pp. 91–123. Westview, Boulder, CO.]

benefit from a large number of activities and technologies. It is particularly instructive to compare perceptions of various radiation and chemical technologies. Nuclear power has a very high perceived risk and low perceived benefit, whereas diagnostic X-rays have the opposite pattern (low perceived risk, high perceived benefit). A parallel finding occurs with chemicals. Nonmedical sources of exposure to chemicals (e.g., pesticides, food additives, alcohol, cigarettes) are seen as very low benefit and high in risk; chemicals used in medicine (e.g., prescription drugs, antibiotics, vaccines) are generally seen as high in benefit and low in risk, despite the fact that they can be very toxic substances. Research throughout the 1980s and 1990s has continued to show high levels of concern regarding the risks from the

use of pesticides. Figures 63.2 and 63.3 present data from parallel national surveys in the United States and France conducted in 1992. The U.S. survey shows significant concerns regarding pesticides in food, which were seen as close in risk to drinking alcohol, motor vehicle accidents, and nuclear power plants and higher in risk than two rather significant hazards, bacteria in food and storms and floods. The picture in France was similar (Figure 63.3). There was a somewhat greater frequency of high-risk responses to pesticides in food in France than in the United States. Similar results were obtained in a 1993 national survey of the Canadian public (Figure 63.4), with more than 30% of Canadians judging pesticides in food as high risk and more than 70% judging them as high or moderate risk.

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Figure 63.2  Perceived health risks to American public: 1992 national survey.

Figure 63.3  Perceived health risks to French public: 1992 national survey.

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Figure 63.4  Health risks to the Canadian public: 1992 Health and Welfare Canada survey (N  1506). [Slovic, P., Malmfors, T., Krewski, D., Mertz, C. K., Neil, N., and Bartlett, S. (1995). Intuitive toxicology. II. Expert and lay judgments of chemical risks in Canada. Risk Anal. 15, 661–675.]

Pesticides in food were rated as higher in risk than drinking alcohol, nuclear power plants, asbestos, and bacteria in food. When the same survey was given to members of the Society of Toxicology of Canada (see Fig. 63.5), pesticides in food were judged far lower in risk (less than 10% judged them as high risk; only about 20% judged them high or moderate risk). Contrary to the public’s opinions, the toxicologists judged bacteria in food to be riskier than pesticides in food. Members of the British Toxicology Society (BTS) surveyed by Slovic et al. (1997) also judged pesticides in food to pose rather small risks (see Fig. 63.6). The most recent data that my colleagues and I have collected come from a national survey in the United States in 1997. Pesticides were seen as posing high risk to the American public by 26% of the respondents and posing high or moderate risk by 69%. Pesticides were rated almost as risky as stored nuclear waste, motor vehicles, nuclear power plants, and natural disasters (see Fig. 63.7). In another segment of that survey, respondents judged risks to individuals to be almost or as great from pesticides as from handguns and violent crime.

63.2.1  Social, Cultural, and Political Influences on Risk Perception The data shown in Figs. 63.1–63.7 reflect only the tip of the iceberg. Below the surface are rumblings of quite complex social, cultural, and political forces shaping the observed ratings of risks from pesticides and other hazards. Recent studies have shown that such factors as gender, race, political worldview, affiliation, emotion, and trust are strongly correlated with risk judgments. Equally important is that these factors influence the judgments of experts as well as the judgments of laypersons. For example, gender is strongly related to risk judgments and attitudes. Several dozen studies have documented the finding that men tend to judge risks as smaller and less problematic than do women (Brody, 1984; Carney, 1971; DeJoy, 1992; Gutteling and Wiegman, 1993; Gwartney-Gibbs and Lach, 1991; Pillisuk and Acredolo, 1988; Sjöberg and Drottz-Sjöberg, 1993; Slovic et al., 1989, 1993; Spigner et al., 1993; Steger and Witt, 1989; Stern et al., 1993). A number of hypotheses have been put

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FIGURE 63.5 Health risks to the Canadian public: 1993 survey of the Society of Toxicology of Canada (N  150). [Slovic, P., Malmfors, T., Krewski, D., Mertz, C. K., Neil, N., and Bartlett, S. (1995). Intuitive toxicology. II. Expert and lay judgments of chemical risks in Canada. Risk Anal. 15, 661–675.]

forward to explain sex differences in risk perception. One approach has been to focus on biological and social factors. Women have been characterized as more concerned about human health and safety because they are socialized to nurture and maintain life (Steger and Witt, 1989). They have been characterized as physically more vulnerable to violence, such as rape, and this may sensitize them to other risks (Baumer, 1978; Riger et al., 1978). The combination of biology and social experience has been put forward as the source of a “different voice” that is distinct to women (Gilligan, 1982; Merchant, 1980). A lack of knowledge and familiarity with science and technology has also been suggested as a basis for these differences, particularly with regard to nuclear and chemical hazards. Women have been discouraged from studying science and there are relatively few women scientists and engineers (Alper, 1993). However, Barke, Jenkins-Smith, and Slovic have found that female physical scientists judge risks from nuclear technologies to be higher than do male physical scientists (Gilligan, 1982; Merchant, 1980). Similar results with scientists were obtained by Slovic et al. (1997), who found that female members of the BTS were far more likely than male toxicologists to judge societal

risks, including pesticides, as moderate or high. Certainly female scientists in these studies cannot be accused of lacking knowledge and technological literacy. Some other factors must influence their decisions. Hints about the origin of these sex differences come from a study by Flynn et al., in which 1512 Americans were asked, for each of 25 hazard items, to indicate whether the hazard posed (1) little or no risk, (2) slight risk, (3) moderate risk, or (4) high risk to society (Flynn et al., 1994). Figure 63.8 shows the difference in the percentage of males and females who rated a hazard as a “high risk.” All differences are to the right of the 0% mark, indicating that the percentage of high-risk responses was greater for women on every item (Flynn et al., 1994). Perhaps the most striking result from this study is shown in Fig. 63.9, which presents the mean risk ratings separately for white males, white females, nonwhite males, and nonwhite females. Across the 25 hazards, white males produced risk-perception ratings that were consistently much lower than the means of the other three groups (Flynn et al., 1994). When the data underlying Fig. 63.9 were examined more closely, Flynn et al. observed that not all white males

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FIGURE 63.6 Health risks to the average exposed citizen of your country: 1994 British Toxicology Society (N  312). [Slovic, P., Malmfors, T, Mertz, C. K., Neil, N., and Purchase, I. F. H. (1997). Evaluating chemical risks: Results of the British Toxicology Society. Hum. Exp. Toxicol. 16, 289–304.]

FIGURE 63.7 Perceived health risks to American public: U.S. population as a whole. [1997 National Risk Survey.]

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Base % Stress 45.3 Suntanning 34.3 Nuclear Waste 52.4 Nuclear Power Plants 25.9 Ozone Depletion 38.8 AIDS 54.7 Drinking Alcohol 34.7 Hi-Volt Power Lines 15.8 Street Drugs 55.6 Motor Vehicle Accidents 32.9 Blood Transfusions 25.1 Chemical Pollution 41.6 Pesticides in Food 32.0 Bacteria in Food 18.7 Cigarette Smoking 57.9 Storms and Floods 11.5 Radon in Home 12.2 Climate Change 22.9 Food Irradiation 18.0 Outdoor Air Quality 24.8 Coal/Oil Burning Plants 18.5 Genet Engr Bacteria 15.0 Medical X-Rays 5.8 Commercial Air Travel 7.3 VDTs 9.6 –10% –5%

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FIGURE 63.8 Perceived health risks to American public by gender: difference between males and females. Note: Base percentage equals male highrisk response. Percentage difference is female highrisk response minus male high-risk response. [Flynn, J., Slovic, P., and Mertz, C. K. (1994). Gender, race, and perception of environmental health risks. Risk Anal. 14, 1101–1108.]

0%

5%

10%

15%

20%

25%

Percent difference in high risk

White male Non-white male

White female Non-white female

Cigarette smoking Street drugs AIDS Stress Chemical pollution Nuclear waste Motor vehicle accidents Drinking alcohol Suntanning Ozone depletion Pesticides in food Outdoor air quality Blood transfusions Coal/Oil burning plants Climate change Bacteria in food Nuclear power plants Food irradiation Storms and floods Genet engr bacteria Radon in home Hi-volt power lines VDTs Medical X-rays Commercial air travel 2 Slight risk

3 Moderate risk

4 High risk

FIGURE 63.9 Mean risk-perception ratings by race and gender. [Flynn, J., Slovic, P., and Mertz, C. K. (1994). Gender, race, and perception of environmental health risks. Risk Anal. 14, 1104.]

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perceived risks as low. The “white-male effect” appeared to be caused by about 30% of the white-male sample who judged risks to be extremely low (Flynn et al., 1994). The remaining white males were not much different from the other subgroups with regard to perceived risk. Further analyses showed that the subgroup of white males who perceive risks to be quite low can be characterized by very great trust in institutions and authorities and by anti-egalitarian attitudes, including a disinclination toward giving decision-making power to citizens in areas of risk management. The results of this study raise new questions. What does it mean for the explanations of gender differences when we see that the sizable differences between white males and white females do not exist for nonwhite males and nonwhite females? Why do a substantial percentage of white males see the world as so much less risky than everyone else sees it? Obviously, the salience of biology is reduced by these data on risk perception and race. Biological factors should apply to nonwhite men and women as well as to white men and women. The present data thus move us away from biology and toward sociopolitical explanations. Perhaps white males see less risk in the world because they create, manage, control, and benefit from many of the major technologies and activities. Perhaps women and nonwhite men see the world as more dangerous because in many ways they are more vulnerable, because they benefit less from many of its technologies and institutions, and because they have less power and control over what happens in their communities and their lives. Although the survey conducted by Flynn et al. was not designed to test these alternative explanations, the race and gender differences in perceptions and attitudes point toward the role of power, status, alienation, trust, perceived government responsiveness, and other sociopolitical factors in determining perception and acceptance of risks. According to this view, the problem of risk conflict and controversy goes beyond science. It is deeply rooted in the social and political fabric of our society.

63.3  Intuitive toxicology: expert and lay judgments of chemical risks The preceding sections have described general attitudes and perceptions regarding pesticides and other hazards. In parallel with this work, another stream of research has focused on chemical risks and attempted to go beneath the surface differences between experts and laypersons to document and understand the causes of these different views. This research has centered around a concept labeled intuitive toxicology (Kraus et al., 1992; Slovic et al., 1995, 1997). Research on intuitive toxicology was motivated by the premise that different assumptions, conceptions, and values

Hayes’ Handbook of Pesticide Toxicology

underlie much of the discrepancy between expert and lay views of chemical risks. Research attempted to address this issue by exploring the cognitive models, assumptions, and inference methods that comprise laypeople’s “intuitive toxicological theories” and comparing these theories with the cognitive models, assumptions, and inference methods of scientists working in the field of toxicology. The work began by identifying several fundamental principles and judgmental components within the science of risk assessment. Questions were developed based on these fundamentals in order to determine the extent to which laypeople and experts share the same beliefs and conceptual framework. Questions addressed the following topics: (a) dose-response sensitivity; (b) trust in animal and bacterial studies; (c) attitudes toward chemicals; (d) attitudes toward reducing chemical risks; (e) conceptions of toxicity, including the toxicity of natural versus synthetic substances and the toxicity of prescription drugs versus chemicals in general; and (f) interpretation of evidence regarding cause-effect relationships between exposure to chemicals and human health. Questions on these topics were incorporated into a survey designed for both experts and the public. Each question was designed, whenever possible, according to a guiding hypothesis about how experts and “lay toxicologists” might respond. For example, a key principle in toxicology is the fact that “the dose makes the poison.” Any substance can cause a toxic effect if the dose is great enough. Thus, experts were predicted to be quite sensitive to considerations of exposure and dose when responding to questions on this topic. In contrast, the often-observed concerns of the public regarding very small exposures or doses of chemicals led to the hypothesis that the public would have more of an “all-or-none” view of toxicity and would be rather insensitive to concentration, dose, and exposure (thus equating any exposure with harm). Because the science of toxicology and the discipline of risk assessment rely so heavily upon animal studies, experts were predicted to have a more favorable view than laypersons regarding the value of such studies. The prediction that laypersons lack sensitivity to dose-response considerations and thus fear even small exposures to toxic or carcinogenic substances led to the prediction that they would exhibit far more negative attitudes toward chemicals than experts. This last prediction was confirmed dramatically in the studies. The members of the public who responded to these surveys associated exposure to chemicals to a remarkable extent with danger, cancer, and death, consistent with the general opinions described in Figs. 63.1–63.7 for pesticides and other chemicals. Specifically, studies of intuitive toxicology on national populations in the United States, Canada, and France have found that about 70% of the public believe that “if a person is exposed to a chemical that can cause cancer, then that person will probably get cancer some day” (Krause

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et al., 1992; Krewski et al., 1995). About 75% of the respondents in these surveys agreed that “If even a tiny amount of a cancer-producing substance was found in my tap water, I wouldn’t drink it.” More than 50% agreed that “There is no safe level of exposure to a cancer-causing chemical.” The concern that any exposure to a carcinogen, no matter how small, is likely to cause cancer is linked to a desire to avoid chemicals and reduce the risks of exposure to them at any cost. About 75% of the public surveyed agreed that “I try hard to avoid contact with chemicals and chemical products in my daily life.” About 62% agreed that “It can never be too expensive to reduce the risks from chemicals.” Responses of toxicologists were not at all in agreement with these views. Of particular importance in this research is the finding, as predicted, that the public is much less sensitive than the experts to considerations of dose and exposure. Although the public recognizes the importance of these factors in some domains (e.g., prescription drugs), they generally tend to view chemicals as either safe or dangerous and they appear to equate even small exposures to toxic or carcinogenic chemicals with almost certain harm. This orientation was found to be associated with high levels of concern regarding chemicals, including very small residues of chemicals on food, and a desire to reduce chemical risks at any cost. Views on the validity of animal studies have been found to be more complex than expected. Consider two survey items that have been studied repeatedly. One is statement S1: “Would you agree or disagree that the way an animal reacts to a chemical is a reliable predictor of how a human would react to it?” The second statement, S2, is a little more specific: “If a scientific study produces evidence that a chemical causes cancer in animals, then we can be reasonably sure that the chemical will cause cancer in humans.” When members of the American and Canadian public responded to these items, they showed moderate agreement with S1; about half the people agreed and half disagreed that animal tests were reliable predictors of human reactions to chemicals. However, in response to S2, which stated that the animal study found evidence of cancer, there was a jump in agreement to about 70% among both male and female respondents (see Fig. 63.10). The important point about the pattern of response is that agreement was higher on the second item. What happens when toxicologists are asked about these two statements? Figure 63.11 shows that toxicologists in the United States and toxicologists in the United Kingdom responded similarly to the public on the first statement but differently on the second (Kraus et al., 1992). They exhibited the same rather middling level of agreement with the general statement about animal studies as predictors of human health effects.1 However, when these studies were said to find evidence of carcinogenicity in animals, the 1 This is a rather surprising result, given the heavy reliance on animal studies in toxicology.

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Figure 63.10  Agreement among members of the U.S. public with statements S1 and S2. [Kraus, N. N., Malmfors, T., and Slovic, P. (1992). Intuitive toxicology: Expert and lay judgments of chemical risks. Risk Anal. 12, 215–232.]

Figure 63.11  Agreement among the public and toxicologists with statements S1 and S2. [Slovic, P. (1997). Trust, emotion, sex, politics, and science: Surveying the risk-assessment battlefield. In “Environment, Ethics, and Behavior” (M. H. Bazerman, D. M. Messick, A. E. Tenbrunsel, and K. A. Wade-Benzoni, eds.), p. 299. New Lexington, San Francisco.]

toxicologists were less likely to agree that the results could be extrapolated to humans. Thus, findings that lead toxicologists to be less willing to generalize to humans lead the public to see the chemical as more dangerous for humans. Figure 63.12 presents the responses for S1 and S2 among men and women toxicologists in the United Kingdom (208 men and 92 women). Here, we see another interesting finding. The men agree less on the second statement than on the first, but the women agree more, just like the general public. Among toxicologists, women are more willing than men to say that one can generalize to humans from positive carcinogenicity findings in animals. These studies of intuitive toxicology have yielded a number of intriguing findings that likely pertain to views about pesticides. One is the low percentage of agreement that animal studies can predict human health effects. Another is that toxicologists show even less confidence in equating human cancers with studies that find cancer in animals resulting from chemical exposure. The public, on

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Figure 63.12  Agreement among men and women toxicologists in the United Kingdom with statements S1 and S2. [Slovic, P. (1997). Trust, emotion, sex, politics, and science: Surveying the risk-assessment battlefield. In “Environment, Ethics, and Behavior” (M. H. Bazerman, D. M. Messick, A. E. Tenbrunsel, and K. A. Wade-Benzoni, eds.), pp. 299. New Lexington, San Francisco.]

the other hand, has high confidence in animal studies that find cancer. These studies also help us understand why the public has come to fear pesticides and other chemicals so greatly. As regulators have sought to develop more effective ways to meet public demands for a safer and healthier environment, they have come to rely heavily on quantitative risk assessment based on animal tests. Such tests often find evidence of cancer at high dose levels. Many scientists are skeptical of such evidence, on the grounds that high doses overwhelm the animals’ defense mechanisms and produce cancers that would not occur in humans under normal conditions of exposure. This skepticism is seen in the high percentage of toxicologists who lack confidence in evidence for carcinogenicity derived from animal studies. The public, on the other hand, exhibits a high degree of confidence in positive findings from animal studies. Thus, the large number of animal studies performed over the years may have done a better job of scaring the public than of informing science about chemical carcinogenesis. Another contributing factor is that interpretation of the animal data has been based on a linear model that assumes that there is no level of exposure to a carcinogen that is without some degree of risk. Multiplying even very small probabilities of contracting cancer across large numbers of exposed individuals will likely project at least some number of deaths. This frightens people. Using upper 95% confidence bounds in the linear extrapolation makes the scenario even more frightening. Thus, the many people who believe there is no safe level of exposure to a carcinogen may have learned this from hearing about the linearity assumption or seeing risk estimates projected from a linear model. Psychological and anthropological research also helps us understand the nature of the public’s fear of exposure to toxic substances that are said (by scientists using a linear

Hayes’ Handbook of Pesticide Toxicology

model) to be toxic at all levels. For example, Frazer (1959) and Mauss (1972) describe a belief, widespread in many cultures, that things that have been in contact with each other may influence each other through transfer of some of their properties via an “essence.” Thus, “once in contact, always in contact,” even if that contact (exposure) is brief. Rozin et al. (1986) show that this belief system, which they refer to as a “law of contagion,” is common in our present culture. The implication of these notions is that even a minute amount of a toxic substance in one’s food (e.g., a pesticide residue) will be seen as imparting toxicity to the food; any amount of a carcinogenic substance will impart carcinogenicity, etc. The “essence of harm” that is contagious is typically referred to as contamination. Being contaminated clearly has an all-or-nothing quality to it—like being alive or pregnant. When a young child drops a sucker on the floor, the brief contact with “dirt” may be seen as contaminating the candy, causing the parent to throw it away rather than washing it off and returning it to the child’s mouth. This all-or-nothing quality, irrespective of the degree of exposure, is evident in the observation by Erikson (1990) that “To be exposed to radiation or other toxins ... is to be contaminated in some deep and lasting way, to feel dirtied, tainted, corrupted” (p. 122). A contagion or contamination model is much more likely to hold in a world in which scientists use linear extrapolation to estimate risks than in a world that recognizes the beneficial effects of chemicals at low doses. We do not, for example, view ourselves as being “contaminated” by exposures to prescription drugs. Another relevant psychological tendency is to confound perception of risk with perception of benefit. If an activity or substance conveys some benefit upon us, we are likely to perceive it as less risky (Alhakami and Slovic, 1994) and more acceptable (Starr, 1969).

Conclusion In this brief overview, an attempt has been made to show the depth and complexity of the public’s concerns regarding the risks from pesticides and other chemicals. These concerns transcend national borders and seem to have held remarkably constant for several decades, in spite of views of many toxicologists and other scientists that are quite the opposite from public views. Fortunately, the research described here will help us understand why public attitudes are the way they are and why they are so resistant to change. These attitudes are a complex product of human psychology and culture interacting with complex and idiosyncratic sciences such as toxicology, epidemiology, and risk assessment. One thing is clear. Risk communication efforts conducted by public relations specialists cannot turn public views around and may, in fact, exacerbate them. Some investigators have taken the limitations of

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risk science, the difficulty of creating and maintaining trust, and the subjective nature of risk judgments as signs pointing to the need for a radically different approach to dealing with conflicts regarding pesticides and other chemical products. This “new” approach focuses on introducing more public participation into both risk assessment and risk decision making to make the decision process more democratic, improve the relevance and quality of technical analysis, and increase the legitimacy and public acceptance of the resulting decisions. Work by scholars and practitioners in Europe and North America has begun to lay the foundation for improved methods of public participation within deliberative decision processes that include negotiation, mediation, oversight committees, and other forms of public involvement (English, 1992; Kunreuther et al., 1993; National Research Council, 1996; Renn et al., 1991, 1995). Those who are concerned about promoting rational decisions about the use of pesticides would be well advised to give careful consideration to this approach.

REFERENCES Alhakami, A. S., and Slovic, P. (1994). A psychological study of the inverse relationship between perceived risk and perceived benefit. Risk Anal. 14, 1085–1096. Alper, J. (1993). The pipeline is leaking women all the way along. Science 260, 409–411. Baumer, T. L. (1978). Research on fear of crime in the United States. Victimology 3, 254–264. Brody, C. J. (1984). Differences by sex in support for nuclear power. Social Forces 63, 209–228. Carney, R. E. (1971). Attitudes toward risk. In “Risk Taking Behavior: Concepts, Methods, and Applications to Smoking and Drug Abuse” (R. E. Carney, ed.). Thomas, Springfield, IL. Carson, R. (1962). “Silent Spring,” Fawcett, New York. DeJoy, D. (1992). An examination of gender differences in traffic accident risk perception. Accident Analysis and Prevention 24, 237–246. English, M. R., (1992). “Siting Low-Level Radioactive Waste Disposal Facilities: The Public Policy Dilemma,” Quorum, New York. Erikson, K. (1990). Toxic reckoning: Business faces a new kind of fear. Harvard Business Rev. 68, 118–126. Flynn, J., Slovic, P., and Mertz, C. K. (1994). Gender, race, and perception of environmental health risks. Risk Anal. 14, 1101–1108. Frazer, J. G. (1959). “The New Golden Bough: A Study in Magic and Religion,” (original work published in 1890). Macmillan Co, New York. Gilligan, C. (1982). “In a Different Voice: Psychological Theory and Women’s Development,” Harvard Univ. Press, Cambridge, MA. Gutteling, J. M., and Wiegman, O. (1993). Gender-specific reactions to environmental hazards in the Netherlands. Sex Roles 28, 433–447. Gwartney-Gibbs, P. A., and Lach, D. H. (1991). Sex differences in attitudes toward nuclear war. J. Peace Res. 28, 161–174. Kraus, N. N., Malmfors, T., and Slovic, P. (1992). Intuitive toxicology: Expert and lay judgments of chemical risks. Risk Anal. 12, 215–232.

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Krewski, D., Slovic, P., Bartlett, S., Flynn, J., and Mertz, C. K. (1995). Health risk perception in Canada. II. Worldviews, attitudes and opinions. Human and Ecological Risk Assessment 1, 231–248. Kunreuther, H., Fitzgerald, K., and Aarts, T. D. (1993). Siting noxious facilities: A test of the facility siting credo. Risk Anal. 13, 301–318. Mauss, M. (1972). “A General Theory of Magic,” (original work published in 1902). Norton, New York. Merchant, C. (1980). “The Death of Nature: Women, Ecology, and the Scientific Revolution,” Harper & Row, New York. National Research Council, Committee on Risk Characterization (1996). “Understanding Risk: Informing Decisions in a Democratic Society” (P. C. Stern and H. V. Fineberg, eds.). Natl. Acad. Press, Washington, DC. Pillisuk, M., and Acredolo, C. (1988). Fear of technological hazards: One concern or many? Social Behavior 3, 17–24. Renn, O., Webler, T., and Johnson, B. B. (1991). Public participation in hazard management: The use of citizen panels in the U.S. Risk— Issues in Health and Safety 2, 197–226. Renn, O., Webler, T., and Wiedemann, P. (1995). “Fairness and Competence in Citizen Participation,” Kluwer Academic, Dordrecht. Riger, S., Gordon, M. T., and LeBailly, R. (1978). Women’s fear of crime: From blaming to restricting the victim. Victimology 3, 274–284. Rozin, P., Millman, L., and Nemeroff, C. (1986). Operation of the laws of sympathetic magic in disgust and other domains. J. Personality and Social Psychology 50, 703–712. Sjöberg, L., and Drottz-Sjöberg, B. M. (1993). “Attitudes Toward Nuclear Waste,” Rhizikon Research Report 12, Center for Risk Research. Stockholm School of Economics, Stockholm. Slovic, P. (1997). Trust, emotion, sex, politics, and science: Surveying the risk-assessment battlefield. In “Environment, Ethics, and Behavior” (M. H. Bazerman, D. M. Messick, A. E. Tenbrunsel, and K. A. WadeBenzoni, eds.), pp. 277–313. New Lexington, San Francisco. Slovic, P., Fischhoff, B., and Lichtenstein, S. (1985). Characterizing perceived risk. In “Perilous Progress: Technology as Hazard” (R. W. Kates, C. Hohenemser, and J. X. Kasperson, eds.), pp. 91–123. Westview, Boulder, CO. Slovic, P., Flynn, J., Mertz, C. K., and Mullican, L. (1993). “Health Risk Perception in Canada,” Report 93-EHD-170. Department of National Health and Welfare, Ottawa. Slovic, P., Kraus, N. N., Lappe, H., Letzel, H., and Malmfors, T. (1989). Risk perception of prescription drugs: Report on a survey in Sweden. Pharm. Med. 4, 43–65. Slovic, P., Malmfors, T., Krewski, D., Mertz, C. K., Neil, N., and Bartlett, S. (1995). Intuitive toxicology. II. Expert and lay judgments of chemical risks in Canada. Risk Anal. 15, 661–675. Slovic, P., Malmfors, T., Mertz, C. K., Neil, N., and Purchase, I. F. H. (1997). Evaluating chemical risks: Results of a survey of the British Toxicology Society. Hum. Exp. Toxicol. 16, 289–304. Spigner, C., Hawkins, W., and Loren, W. (1993). Gender differences in perception of risk associated with alcohol and drug use among college students. Women and Health 20, 87–97. Starr, C. (1969). Social benefit versus technological risk. Science 165, 1232–1238. Steger, M. A. E., and Witt, S. L. (1989). Gender differences in environmental orientations: A comparison of publics and activists in Canada and the U.S. Western Political Quarterly 42, 627–649. Stern, P. C., Dietz, T., and Kalof, L. (1993). Value orientations, gender, and environmental concern. Environ. Behav. 25, 322–348.

Section X

Organophosphorous and N-Methyl Carbamate Insecticides

(c) 2011 Elsevier Inc. All Rights Reserved.

Chapter 64

Chemistry of Organophosphorus Insecticides Howard W. Chambers1, Edward C. Meek2 and Janice E. Chambers2,* 1

Department of Entomology and Plant Pathology, Mississippi State University, Mississippi State, Mississippi, USA 39762. Center for Environmental Health Sciences, College of Veterinary Medicine, Mississippi State University, Mississippi State, Mississippi, USA 39762.

2

The following review is intended to introduce and summarize the chemistry of organophosphorus insecticides, primarily for toxicologists. It is in no way designed to give in depth coverage of the entire subject of organophosphorus chemistry. More extensive reviews may be found in Eto (1961) and Fest and Schmidt (1982).

the safest OPs ever marketed. By 1959, it was estimated that more than 50,000 OP compounds had been made. In 1970, more than 200 OP insecticides were marketed worldwide. Though development of resistance in pests and the marketing of new, safer insecticides have greatly decreased the usage of OPs and slowed the development of new products, these remain an important group of pest control agents and will probably do so for another decade or more.

64.1.1  History

64.1.2  Classification and Nomenclature

Organophosphorus (OP) chemistry apparently began around 1820 with the esterification of alcohols to phosphoric acid. Despite synthesis of a number of OP compounds in the early 1900s, the potential toxicity went unrecognized until the 1930s. By 1940, groups led by B. C. Saunders in England and Gerhard Schrader in Germany had produced several highly toxic compounds for possible use as chemical warfare agents. The most notable of these were sarin and soman, both phosphorofluoridates, and tabun, a phosphorocyanidate (Fig. 64.1). Schrader’s group also produced the first commercial OP insecticides, including tepp in 1937, dimefox in 1940, schradan (OMPA) in 1942, and parathion in 1944. Following World War II, with the capture of Schrader’s research records, interest in OP insecticides grew rapidly. Although all of the early chemicals were effective insecticides, they were also highly toxic to mammals. In 1950, however, American Cyanamid produced malathion, one of

Because of the vast number and wide variety of organic phosphorus chemicals that can exist, any comprehensive classification system would be too complex to undertake here. Instead, a classification scheme will be presented into which all commercially important OP insecticides will fit, based on the central phosphorus atom and the four atoms immediately surrounding it. Similarly, the nomenclature of OPs will primarily address these five atoms. The general structure of OP insecticides can be represented by (Fig. 64.2). L, the so-called leaving group, is the most reactive and most variable substituent. The term “leaving group” comes from the fact that it is the substituent that is displaced when the OP phosphorylates acetylcholinesterase, the primary target enzyme. The leaving group is also usually the most susceptible to hydrolysis. R1 and R2 are less reactive and are most commonly alkoxy groups, but may be alkyl, aryl, alkylthio, or alkylamino. X is either oxygen or sulfur. OP insecticides may be considered to be derivatives of phosphoric acid (H3PO4) or phosphonic acid (H3PO3) in which all H atoms are replaced by organic moieties. Thus, phosphates are compounds in which the P atom is surrounded by four O atoms. In phosphonates, there are three

64.1  Introduction

* College of Veterinary Medicine, 240 Wise Center Drive, Mississippi State University, Mississipi State, MS 39762-6100. Email: [email protected]. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

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CH3

O (CH3)2CHO P

F

P

F

C

N

(CH3)2N

CH3

Sarin Figure 64.1  Structures of early OP nerve gases.

Tabun

Soman

64.1.3  Synthesis

X R2

O CH3CHO

P

CH3

R1

O

(CH3)3CCHO

P

L

Figure 64.2  General structure of OP insecticides.

O atoms and one phosphorus-carbon bond. Phosphinates, which have two O atoms and two P—C bonds, have been investigated and show biological activity, but none have been developed commercially. In many OPs, one or more of the oxygen atoms are replaced by sulfur and/or nitrogen. For phosphoric acid derivatives, the O, S and N atoms can be arranged in 20 different configurations. Another 12 different configurations can exist for derivatives of phosphonic acid. Fortunately for the classification and nomenclature to be considered here, most of the 32 possible configurations have not appeared in commercial OP insecticides. Figure 64.3 shows the eight subclasses and thirteen configurations for which one or more commercial OP insecticides are known. (Fig. 64.3)

Because of the wide variety of organophosphorus compounds known, no attempt will be made here to cover the many different pathways involved in their synthesis. Rather, the preparation of the seven most important intermediates will be presented, followed by representative examples of their use in the synthesis of specific insecticides. For additional information, the reader is referred to more comprehensive reviews by Eto (1961) and Fest and Schmidt (1982). Initially, elemental phosphorus is converted into P2S5 by reaction with sulfur or into PCl3 by direct chlorination. These two materials are then converted into the following seven intermediates from which most OP insecticides are synthesized: S (1)

P2S5 � 4 R

OH

2 (R

S (2) O

O

O

P O O Phosphate O

S

O

Phosphorothioates

O

P O

C

C

Phosphonothioates O

O

O

O

P S

Phosphoramidate

(R

O

(3)

P S

S P S

C

Cl

PCl3 � ½O2

PCl3

O P O

S

S

PCl3 � S

(4)

Phosphonate O

O)2 P

O

C



O P S

PCl3

(5)

PCl3 � 3 R

OH

base

(R

O)3P

(6)

PCl3 � 3 R

OH

no base

(R

O)2P

C

Phosphonodithioates O

P O

N

SH � ½ Cl2

O

O

S

S

P S

P O

Phosphorodithioates O

O)2 P

SH

O

O

O

S

P S

S

(R

O)2P

N

S P O

O

O

OH

P S

N

Phosphoramidothioates

Figure 64.3  Nomenclature of major subclasses of OP insecticides.

O (7)

PCl3 � 2 R

OH � ½O2

(R

O)2P

Cl

Chapter | 64  Chemistry of Organophosphorus Insecticides

Because there are almost as many procedures for formation of the P—C bond of phosphonates and related OPs as there are insecticides containing this bond, the synthesis of required intermediates will not be discussed. It may be noted, however, that most processes involve PCl3, AlCl3, and alkyl halides. Trialklyl phosphites (intermediate 5) are particularly useful in the preparation of dialkyl vinyl phosphates from -chloroaldehydes and ketones. The synthesis of DDVP serves as a good example: CH3O CH3O

O

1397

The free acid will add across certain carbon-carbon double bonds. Malathion is prepared from diethyl maleate by this reaction: S

O

P

SH � CHC

OCH2CH3

CH3O

CHC

OCH2CH3

CH3O

S

CH3O CH3O

P

O S

CHC

OCH2CH3 OCH2CH3

CH2C

O

O

Also, the free acid reacts with formaldehyde and a mercaptan to produce compounds such as phorate:

O CH3O

P � HCCCl3

CH3O

S

P

O

CH

CCl2

CH3CH2O CH3CH2O

P

S SH � CH2O � HS

CH2CH3

CH3CH2O CH3CH2O

P

CH2 S

S

CH2CH3

CH3O

Interestingly, the same aldehyde reacts with dimethyl phosphite (intermediate 6) to produce the phosphonate insecticide trichlorfon: O CH3O

P

CH3O

O CH3O

OH � HCCCl3

OH

P

CH3CH2O

Cl � HO

NO2

P

CH3O

base

CH3CH2O CH3CH2O

CCCl3

P

O

NO2

An analogous, and more commonly used, reaction is that of dialkyl phosphorochloridothioates (intermediate 2) to produce phosphorothionates such as parathion: S CH3CH2O

S

P

CH3CH2O

Cl � HO

NO2

base

CH3CH2O CH3CH2O

P

O

NO2

Dialkyl phosphorodithioates undergo three distinct types of reactions. The salts of these acids react with alkyl halides with formation of the thiol ester. Disulfoton is produced in this way: S CH3CH2O CH3CH2O

P

S S

Na � Cl

CH2CH2SCH2CH3

CH3CH2O CH3CH2O

P

CH3

O

Dialkyl phosphates may also be prepared from dialkyl phosphorochloridates (intermediate 7) by reactions analogous to the formation of carboxylic esters from acid chlorides. Paraoxon, the oxygen analog and active metabolite of parathion, is readily prepared by this reaction: O O CH3CH2O

For OPs with three different substituents on the P atom, it is necessary to begin with P(:O) Cl3 (from intermediate 3) or P(:S) Cl3 (from intermediate 4). The sequence in which the groups are added varies with the specific compound. Two examples, fenamiphos and sulprofos, respectively, are as follows:

S

CH2CH2SCH2CH3



S

PCl3 � (1) HO � (2) CH3CH2

OH

� (3) CH3CH

NH2

CH3 CH3

O CH3CH2O

P

S

O

CH3

CH3CHNH

CH3

CH3 S



PCl3 � (1) CH3CH2

OH

� (2) HO

S

CH3

� (3) CH3CH2CH2

S

Na

S CH3CH2O

P

O

NO2

CH3CH2CH2S

Although the preceding discussion is by no means a comprehensive treatment of OP synthesis, the reactions shown are used in the preparation of more than 90% of commercial OP insecticide compounds. One final type of reaction not previously mentioned is worthy of consideration because of its usefulness in laboratory syntheses. Dialkyl phosphates with phenolic or heterocyclic leaving groups are easily prepared from the phenol or heterocyclic alcohol and the dialkyl phosphorochloridate. Unfortunately, the only phosphorochloridates readily available commercially are the dimethyl and the diethyl, the former being quite unstable. At least four dialkyl phosphites (dimethyl, diethyl, di-i-propyl, and di-n-butyl) are available at reasonable prices and are more stable in storage. The reaction alluded to in the previous paragraph, then, is the synthesis of dialkyl phosphates from dialkyl phosphites without independent preparation and isolation of the phosphorochloridate.

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Carbon tetrachloride, in the presence of an organic base (e.g., triethylamine), will chlorinate dialkyl phosphites by the following reaction: O (R

O)2 P

O H � CCl4

(CH3CH2)3N

(R

O)2 P

Cl � CHCl3

This allows a one-step synthesis of many phenyl or heterocyclic dialkyl phosphates. For example, methyl paraoxon (catalog price  $500/g) is prepared easily and inexpensively by: O CH3O CH3O

P

O H � HO

NO2

(CH3CH2)3N

CH3O

in CCl4

CH3O

P

O

NO2

This synthesis has been done several times by the authors, obtaining moderate yield of product of 98% purity. Although the process is useful on a small scale, it is rarely used in industrial syntheses because of the hazard associated with CCl4.

64.1.4  Reactions OP insecticides, when kept cool, dark, and anhydrous, are usually quite stable. Exposure to heat, light (especially ultraviolet), and/or water, however, may lead to chemical alterations. The three primary reactions involving the phosphorus atom and those immediately surrounding it are hydrolysis, oxidation, and rearrangement. Except at very low pHs, hydrolysis of the P—O—C linkage results primarily by OH—attack on the P atom with cleavage of the P—O bond. Thus, rates of hydrolysis increase with increasing pH. Three additional generalizations may be made concerning this type of hydrolysis: 1. Compounds containing PO hydrolyze faster than analogous compounds containing PS. 2. Cleavage occurs between the P atom and the leaving group. 3. The hydrolysis rate decreases with increasing size of the alkyl substituents. The most notable exception is that alkaline hydrolysis of OPs containing Me—O—P(:S) often results in cleavage of the methyl rather than the leaving group. Hydrolysis of the P—S—C linkage differs from that described previously in that alkaline hydrolysis results

primarily in cleavage of the S—C bond. An exception to this is fonofos (O-ethyl S-phenyl ethylphosphonodithioate), which is apparently hydrolyzed at the P—S bond to yield thiophenol. Acid hydrolysis, on the other hand, consistently leads to cleavage of the P—S bond. The P—N bond of phosphoramides is generally rather resistant to alkaline hydrolysis. Because the N atom is readily protonated, however, these compounds are quiet susceptible to acid hydrolysis. Oxidation of OPs by O2 is enhanced by ultraviolet (UV) light. It may also be accomplished by oxidants such as HNO3 or organic peroxyacids. These oxidations most commonly occur with PS-type OPs, but the sulfur of the P—S—C linkage may also be oxidized. Oxidation of PS presumably results in transient formation of the phosphooxythiirane (a three-membered ring consisting of one each of P, O, and S). This intermediate spontaneously decomposes to produce the oxon (PO) with loss of sulfur or cleaves between the P atom and the leaving group. Oxidation of the sulfur of P—S—C produces ester-sulfoxides, which are highly reactive and usually degrade rapidly. Upon exposure to UV or high temperatures, compounds containing C—O—PS will undergo rearrangement in which C—S—P  O is produced. Most commonly, the C involved is in an alkyl substituent but the leaving group can also be involved. Both types of rearrangements are known for parathion as shown by O

CH3CH2O CH3CH2O

CH3CH2S S P

O

NO2

UV

CH3CH2O

P

O

NO2

S

NO2

O CH3CH2O CH3CH2O

P

Whether such rearrangements are intramolecular, intermolecular, or both is unclear. Toxicologically, the chemical reactions may or may not result in loss of toxicity. Hydrolysis completely detoxifies the OP. Oxidation of PS to PO and both rearrangements illustrated lead to an increase in toxicity. Though other reactions of OPs may occur, those presented are the most important in the environment and in long-term storage.

References Eto, E. (1961). “Organophosphorus Pesticides: Organic and Biological Chemistry,” CRC Press, Cleveland. Fest, C., and Schmidt, K.-J. (1982). “The Chemistry of Organic Pesticides,” 2nd ed. Springer-Verlag, Berlin/Heidelberg/New York.

Chapter 65

The Metabolism of Organophosphorus Insecticides Janice E. Chambers1*, Edward C. Meek1 and Howard W. Chambers2 1

Center for Environmental Health Sciences, College of Veterinary Medicine, Mississippi State University, Mississippi State, Mississippi, USA 39762 2 Department of Entomology and Plant Pathology, Mississippi State University, Mississippi State, Mississippi, USA 39762

65.1  Introduction As was described in the previous chapter, the organo­ phosphorus insecticide class contains a diverse array of structures, all united by the presence of a pentavalent phosphorus atom with three singly bonded constituents and a coordinate covalent bond (typically drawn as a dou­ ble bond) to either a sulfur or an oxygen. These insecti­ cides or their metabolites are potent inhibitors of serine esterases through phosphorylation of the serine hydroxyl moiety within the active site of the esterase. The primary target esterase from a toxicological standpoint is acetyl­ cholinesterase, a widely distributed enzyme within the ver­ tebrate nervous system which mediates hydrolysis of the neurotransmitter acetylcholine throughout the central and peripheral nervous systems. The phosphorylation of acetyl­ cholinesterase is relatively persistent, with spontaneous hydrolysis, and therefore recovery of the enzyme activity, requiring hours to days. The inhibition of acetylcholinester­ ase results in the accumulation of acetylcholine in choliner­ gic synapses and neuromuscular/glandular junctions with subsequent hypercholinergic activity. Such activity leads to a variety of signs and symptoms of intoxication, with death in mammals in lethal level poisonings resulting from respi­ ratory failure. Other serine esterases, such as butyrylcholine­ sterase or carboxylesterases, can be phosphorylated by the organophosphorus insecticides or their metabolites also; however, phosphorylation of these other targets does not appear to result in toxic responses. More detailed descrip­ tions of the neurotoxicity of organophosphorus insecticides can be found in a number of chapters and reviews including Chambers (1992) and Ecobichon (2001).

* College of Veterinary Medicine, 240 Wise Center Drive, Mississippi State University, Mississipi State, MS 39762-6100. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

The potency of the organophosphorus insecticides or their active metabolites as inhibitors of target brain acetylcholinesterase does not correspond to the acute tox­ icity levels, indicating that metabolism and disposition are of great significance in determining the overall acute toxic­ ity level of these insecticides (Chambers et al., 1990). An overview of organophosphorus insecticide chemistry, bio­ chemistry and toxicology can be found in Chambers and Levi (1992). These insecticides display substantial chemical diver­ sity, including a variety of atoms in addition to the carbon and phosphorus required by the compounds being “organo­ phosphorus” compounds, such as sulfur, nitrogen and oxy­ gen. Therefore, the organophosphorus insecticides are subject to many metabolic pathways mediated by several of the groups of xenobiotic metabolizing enzymes. The group connected through the single bond which is the least thermodynamically stable of the three single bonds is the “leaving group” and is the group which is eliminated from the molecule as it phosphorylates its esterase targets. Since the leaving group frequently is a more complex structure, such as an aromatic or heterocyclic ring, it may be subject to some metabolic pathways that the other substituents are not, as described further below. The other two substituents may be the same as one another or different, and are also subject to metabolism. The organophosphorus insecticides or their metabolites are subject to Phase 1 reactions (oxida­ tions, reductions, hydrolyses) and Phase 2 reactions (conju­ gations). Because of their metabolic and chemical lability, they do not readily remain intact either in the environment or in the organism. Their environmental lability was one of the factors which allowed them to replace the highly stable organochlorine insecticides as the dominant class of insecticides. An overview of the several types of reactions which occur in the metabolic pathways of organophosphorus 1399

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insecticides will be discussed below, indicating the types of enzymes involved in the reaction, some examples of these reactions with specific organophosphorus insecticides, and the toxicological outcomes of these metabolic pathways. An overview of the types of biotransformation enzymes involved in the metabolism of organophosphorus insec­ ticides can be found in Parkinson (2001). Specific meta­ bolic pathways for a number of specific insecticides can be found in Aizawa (1982, 1989) and Dikshith (1991).

65.2  Oxidations 65.2.1  Cytochromes P450 The cytochrome P450 enzymes (CYP’s; P450’s) com­ prise, in general, one of the most important, if not the most important, class of xenobiotic metabolizing enzymes. The CYP’s are a superfamily of related enzymes character­ ized by the presence of a heme iron in the active site. They have a broad substrate specificity, with different enzymes (isoforms) catalyzing a variety of oxidations on phospho­ rus, sulfur, nitrogen and carbon, with different substrate specificities among the isoforms. The CYP’s are mono­ oxygenases and catalyze the oxidations by the addition of one atom of molecular oxygen into the substrate via an electron transport pathway with the electrons supplied by NADPH and sometimes NADH. The electron transfers are catalyzed by NADPH cytochrome P450 reductase. Except for some specialized CYP’s, such as those involved in steroidogenesis, the cytochrome P450 pathways occur in the endoplasmic reticulum of vertebrate cells, with the high­ est xenobiotic metabolizing capacity in the mammalian liver, and a more limited capacity observed in other mam­ malian tissues and in sub-mammalian vertebrates. More detailed descriptions of cytochrome P450 and associated reactions may be found in Parkinson (2001). Many of the CYP-mediated reactions on organophosphorus insecticides are obvious oxidations, resulting in a more highly oxidized product with the presence of the oxygen apparent in the products. However, some of the CYP-mediated reactions are not as obviously oxidations. Because of the addition of polar reactive groups by these CYP-mediated reactions, some of the resultant products are more biologically reac­ tive and therefore more toxic (such as with greater reac­ tivity toward neural target molecules or toward DNA) than the parent compounds were, while other reactions result in

CH3CH2O

S P

detoxified products; therefore, CYP’s mediate both bioacti­ vations and detoxications. One of the most important of the CYP-mediated reac­ tions involving the organophosphorus insecticides is the desulfuration reaction occurring with phosphorothion­ ates and other compounds having the phosphorus bonded to sulfur by a coordinate covalent bond (typically drawn as PS). The desulfuration reaction involves an attack of the phosphorus by oxygen, resulting in a putative phospho­ oxythiiran intermediate, which rearranges to a PO group with a loss of the sulfur as illustrated in Fig. 65.1 (Neal, 1980). The sulfur released in the desulfuration reaction is a reactive moiety, and has the ability to destroy surround­ ing biomolecules, such as the CYP’s themselves. A clas­ sic example of this desulfuration reaction is the conversion of parathion to its phosphate (oxon) metabolite, paraoxon. The PS compounds are relatively poor anticholinester­ ases, whereas the oxons are potent anticholinesterases, with a three order of magnitude difference in potency with at least some of the insecticides (Forsyth and Chambers, 1989). Since so many of the most popular of the organo­ phosphorus insecticides are phosphorothionates or related PS compounds, the desulfuration reaction is required for them to display appreciable anticholinesterase activity, and therefore to display classical organophosphorus insecticide neurotoxicity. Because so many of the phosphorothion­ ate insecticides display very high acute toxicity levels (for example, rat oral LD50’s for parathion, methyl parathion and azinphosmethyl are 2, 50, and 10 mg/kg, respectively; Meister, 2004), one can infer that the desulfuration reaction occurs in vivo to an appreciable extent. An illustration of the desulfuration of parathion is given in Fig. 65.2. A reaction occurring parallel to the desulfuration reac­ tion, and concurrently with it, is the dearylation reaction, occasionally termed oxidative hydrolysis, which occurs from the same putative phosphooxythiiran intermediate described above for the desulfuration reaction. Instead of rearrang­ ing to eliminate the sulfur as occurs during the desulfura­ tion reaction, the rearrangement in the dearylation reaction

R1

P R2

½O2

S R3

O

S

R2

NO2

Figure 65.2  CYP-mediated desulfuration of parathion to paraoxon.

P

R1 O

R3

P R2

O P

O

CH3CH2O Parathion

O

Figure 65.1  General CYP-mediated desulfuration reaction.

CH3CH2O

CH3CH2O

R1

Paraoxon

NO2

R3 + S:

Chapter | 65  The Metabolism of Organophosphorus Insecticides

eliminates the aryl leaving group. The resultant products are the leaving group plus either the dialkyl phosphorothioate or the dialkyl phosphate; therefore, the reaction resembles a hydrolysis, but the occurrence of these reaction products are dependent upon the presence of cytochrome P450, oxygen and NADPH. Additionally, classic hydrolysis reactions do not readily occur with the phosphorothionates. Therefore, the dearylation reaction appears to be an oxidation although the products do not readily suggest that an oxidation has occurred. Because the phosphate/oxon structure is required for anticholinesterase activity, the dearylation reaction is a detoxication reaction. The dearylation reaction with para­ thion as an example is shown in Fig. 65.3. The concurrent and competing reactions of desulfuration (activation) and dearylation (detoxication), again using para­ thion as an example, are illustrated in Fig. 65.4, along with the putative phosphooxythiiran intermediate. Studies conducted with purified isoforms of the CYP’s have indicated that differ­ ent isoforms have different desulfuration to dearylation ratios, indicating that substrate specificity and pathway preference among the CYP isoforms differ (Levi et al., 1988). Therefore, it is expected that the activity of different isoforms of CYP’s would have different impacts on toxicity. An additional CYP-mediated reaction which can occur on the intact phosphorothionate or its oxon is a S

CH3CH2O

1401

dealkylation reaction in which one of the carbons in an alk­ oxy group is oxidized to the aldehyde which is removed, leaving a hydroxyl group associated with the phosphorus (Appleton and Nakatsugawa, 1972). The oxidized product would be formaldehyde in the case of a methoxy group or acetaldehyde in the case of an ethoxy group. Using para­ thion as an example once again, the dealkylation reaction is illustrated in Fig. 65.5. Oxidations of substituents in the leaving group are also possible, with a wide variety of reactions possible because of the great diversity of the leaving groups within the insecticide class. A few illustrative examples are provided in Fig. 65.6.

65.2.2  Flavin Monooxygenases An additional class of monooxygenases capable of oxidiz­ ing N, P or S occurring in xenobiotics are the flavin mono­ oxygenases (FMO’s) which are also microsomal, most prevalent in the mammalian liver, insert one atom of molec­ ular oxygen into the substrate molecule and require NADPH (Levi and Hodgson, 1992). They have a flavin group instead of a heme group to catalyze the substrate oxidations. There are fewer isoforms of the FMO’s than the CYP’s, and they have more limited substrate specificity than the CYP’s. One CH3CH2O

P

NO2

O

S(O) P

CH3CH2O

OH

+

NO2

HO

CH3CH2O Parathion

Diethyl phosphorothioic acid or diethyl phosphate

4-nitrophenol

Figure 65.3  CYP-mediated dearylation of parathion.

CH3CH2O

CH3CH2O

S P

O P

NO2

O

S

CH3CH2O

O

CH3CH2O

P

NO2

O

CH3CH2O

CH3CH2O NO2

O

S(O) P

CH3CH2O

OH

+

NO2

HO

CH3CH2O

Figure 65.4  CYP-mediated metabolism of parathion through a putative phosphooxythiiran intermediate.

CH3CH2O

S P

O O

NO2

CH3CH

CH3CH2O

HO +

S P

O

CH3CH2O Parathion

Figure 65.5  CYP-mediated deethylation of parathion.

Acetaldehyde

Desethyl parathion

NO2

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1402

of the more important reactions catalyzed by the FMO’s is the sulfoxidation of phorate to the sulfoxide then to the sul­ fone (Fig. 65.7). The sulfoxidaton of the oxon of phorate to its sulfoxide, then to its sulfone, is also possible.

environmentally because of the greater opportunity to provide reducing environments. Using, once again, parathion as an example, the nitro group on the aromatic ring of the leaving group can be reduced, yielding amino parathion (Fig. 65.8).

65.3  Reductions

65.4  Hydrolysis

Reductions are possible outcomes of CYP-mediated reactions, though these would be considered rare in mammalian systems which are oxidizing environments with few exceptions (ex., gut contents). Nevertheless, reductive reaction products have been discovered, and might be expected to occur to a limited extent. Reduction reactions are probably of greater significance

Catalytic hydrolysis of the phosphates/oxons with elimi­ nation of the leaving group is catalyzed by the A-esterases (phosphotriesterases), which are hydrolases designated as capable of hydrolyzing organophosphates and are not inhibited by them (Aldridge, 1953). In the mammalian system with the insecticidal compounds or oxons, the

CH3O

O P

O O

C

CH3O

C

CH2CH3

C

CH3O

N CH2CH3

CH3 Cl

O

O

P

O

C

CH3O

C

CH2CH2OH

C

N CH2CH3

CH3 Cl

Phosphamidon

CH3O

S P

O

CH3O

NO2

CH3O

CH3O

CH3

S P

O

S

CH3O

NO2

P

O

NO2

CH3O

CH2OH

COH

Fenitrothion

CH3CH2O

O

S P

CH(CH3)2

N O

N

CH3CH2O

CH3

Diazinon

S

CH3O

P

O S

CH3O

C

CH2

NHCH3

S

CH3CH2O

CH(CH3)2

N

P

O

N

CH3CH2O

CH3O

CH2OH

S P

O S

CH3O

CH2

C

NHCH2OH

Dimethoate Figure 65.6  Examples of CYP-mediated oxidations.

CH3CH2O CH3CH2O

S P

S

CH2

S

CH2CH3

CH3CH2O

S P

O S

CH3CH2O

CH2

S

CH2CH3

Phorate

CH3CH2O CH3CH2O Figure 65.7  FMO-mediated oxidation of phorate.

S P

O S

CH2

S O

CH2CH3

Chapter | 65  The Metabolism of Organophosphorus Insecticides

A-esterases are calcium dependent. Other A-esterases have greater specificity for diisopropylfluorophosphate and some of the nerve agent phosphates, and have different metal cofactor requirements. The A-esterases are largely micro­ somal, and do not have the diversity of isoforms as the oxi­ dative enzymes. Like the oxidation enzymes, they occur in the highest activity levels in the liver of the mammal, and at lesser activity levels in extrahepatic tissues. Phosphate/ oxon hydrolysis is a detoxication reaction. The A-esterases have a relatively high affinity for a few phosphates/oxons, such as chlorpyrifos-oxon and diazoxon, the active metabo­ lites of chlorpyrifos and diazinon, respectively, but only a very low affinity for many, perhaps most, of the phos­ phates/oxons (Chambers et al., 1994, Furlong et al., 1989, Pond et al., 1996, 1998). Therefore, the in vivo importance of the A-esterases is probably great for a few insecticides, but is difficult to estimate for many of the insecticides. The A-esterases do not appear to hydrolyze PS compounds. Even though a relatively poor substitute, for the sake of consistency throughout this article, the A-esterase-mediated hydrolysis of paraoxon, the active metabolite of parathion, is illustrated in Fig. 65.9. Even though paraoxon is a poor substrate, the hydrolysis of paraoxon has frequently been used to describe and assay A-esterase activity, so the term paraoxonase is often used for this activity, regardless of S

CH3CH2O

1403

the substrate. The abbreviation PON has become popular for paraoxonase activity. PON has been shown to hydro­ lyze oxidized phospholipids and is important in protecting against cardiovascular disease, an activity which probably is PON’s physiological role (Chambers, 2008). Non-catalytic hydrolysis of the phosphates/oxons also occurs when these compounds phosphorylate serine ester­ ases, such as carboxylesterases, butyrylcholinesterase and even the target acetylcholinesterase; all of these esterases are classified as B-esterases, hydrolases which are inhib­ ited by organophosphates and which cannot catalytically hydrolyze them (Aldridge, 1953). These reactions would not be considered metabolism, because the phosphoryla­ tion is persistent, leading to a stoichiometric destruction of one phosphate/oxon molecule per serine esterase molecule, with enzyme incapacitation occurring for a long period of time. Nevertheless, the phosphorylation event releases the leaving group of the molecule, which is the same product produced in dearylation and catalytic hydrolysis reactions. Therefore, serine esterase phosphorylation con­ tributes leaving group to the pool of metabolite formed by catalytic reactions, and this amount is sufficiently high in in vitro preparations to be conveniently measured (Tang and Chambers, 1999). A schematic of the phosphorylation of a serine esterase by paraoxon is given in Fig. 65.10. O

CH3CH2O

P

NO2

O

CH3CH2O

P

NH2

O

CH3CH2O Parathion

Aminoparathion

Figure 65.8  Reduction of parathion. CH3CH2O

O P

H2O

CH3CH2O

NO2

O

O P

CH3CH2O

OH

+

NO2

HO

CH3CH2O Paraoxon

Diethyl phosphate

4-nitrophenol

Figure 65.9  A-esterase mediated hydrolysis of paraoxon. Serine Protein

CH3CH2O

OH

CH3CH2O

O P

CH3CH2O

O

Serine esterase

Phosphorylated esterase

+

+

O P

O

NO2

NO2

HO

CH3CH2O Paraoxon Figure 65.10  Phosphorylation of a serine esterase by paraoxon.

4-nitrophenol

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O CH3O

S P

S

CH

CH3O

O

CH2COCH2CH3

CH3O

COCH2CH3

CH3O

S P

S

CH

CH2COCH2CH3 COH

O

O

Malathion

O S

CH3O

P

S

CH2COH

CH

CH3O

COH O O

CH3O

S P

S

CH

CH3O

CH2COH COCH2CH3 O

Figure 65.11  Carboxylesterase-mediated hydrolysis of malathion.

The carboxylesterases also perform a very important catalytic hydrolysis of the carboxylic acid esters in mala­ thion and contribute greatly to the low mammalian toxicity of malathion. Hydrolyses to the - and -monoacids and the diacid occur from the parent malathion (Fig. 65.11). In mammals these detoxifying hydrolyses occur more readily than the CYP-mediated desulfuration, allowing the mala­ thion to be effectively detoxified prior to appreciable bio­ activation, and resulting in a very low acute toxicity level (rat oral LD50 of 1200 mg/kg; Meister, 1990). Some representative metabolic schemes for a few important organophosphorus insecticides illustrating the major oxidations and hydrolyses are given in Figs. 65.12 and 65.13.

65.5  Conjugations The Phase 2 (conjugation) reactions can render the insec­ ticides or metabolites even more water soluble. These conjugations frequently occur with the leaving groups produced by organophosphate hydrolysis, such as some of the phenols and heterocyclic alcohols or amines. These hydrolytic metabolites would already be detoxified prod­ ucts, at least with respect to anticholinesterase activity, so further metabolism would not be necessary for additional detoxication. Therefore, the Phase 2 reactions have far less impact on toxicity than the Phase 1 reactions do. However, these conjugation reactions will render the metabolites more water soluble than the parent compound or the inter­ mediate metabolites, and therefore allow the metabolites to

be readily excreted. Sulfate and glucuronide conjugates are possible, catalyzed by sulfotransferases and glucuronosyl transferases, respectively; both types of conjugates are hydrophilic. Glutathione conjugation is also possible. Glutathione transferases can theoretically mediate the dealkylation, primarily with methoxy compounds, of organophospho­ rus insecticides, such as the demethylation of methyl para­ thion. However, the in vivo significance of this reaction is controversial (Sultatos, 1992).

65.6  Summary The organophosphorus insecticides are metabolically highly labile, as illustrated by the above discussion. This metabolic lability, along with their general lack of extreme lipophic­ ity, prevent their bioaccumulation. A variety of oxidation, reduction, hydrolysis and conjugation reactions are pos­ sible within the group of organophosphorus insecticides. The mechanism of their acute toxicity is the inhibition of acetylcholinesterase. Some of the organophosphorus insec­ ticides are active anticholinesterases, and any metabolism is therefore a detoxication. Many of the insecticides, however, are not active anticholinesterases in their parent form, and require bioactivation in order to be effective anticholinester­ ases. The CYP-mediated desulfuration reaction is responsi­ ble for the majority of these bioactivations. Most other routes of metabolism would be detoxications. The fact that many of the insecticides or the active metabolites of those insec­ ticides requiring bioactivation are potent anticholinesterases

Chapter | 65  The Metabolism of Organophosphorus Insecticides

CH3CH2O

S

CH(CH3)2

N

P

1405

O

CH3CH2O

N

CH3CH2O

S P

N O

N

CH3CH2O

CH3

CH(CH3)2

CH2OH

Diazinon S

CH3CH2O

P

COH(CH3)2

N O

N

CH3CH2O

CH3CH2O

O

CH(CH3)2

N

P

CH3

O

N

CH3CH2O

CH3 CH3CH2O

H

O P

N OH

+

O

N

CH3CH2O

CH3 O

O CH3O

S P

S

CH

CH3O

CH(CH3)2

CH2COCH2CH3

CH3O

COCH2CH3

CH3O

S P

S

CH2COCH2CH3

CH

COH O

O

O

Malathion

CH3O

S P

S

CH3O

CH

CH2COH COH O

O CH3O CH3O

O P

S

CH

O

CH2COCH2CH3 COCH2CH3 O

CH3O CH3O

S P

S

CH

CH2COH COCH2CH3 O

Figure 65.12  Metabolism of diazinon and malathion.

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CH3O CH3O

O P

O

C

CH

Cl

CH3O

Cl

HO

O O

P

CH

C

Cl Cl

dichlorvos CH3O CH3O

Cl Cl

O C

Cl

O

P

P

OH

+

HC

OCH3

Cl

Cl

CH

Cl

Cl

O

O + HO

C

OCH3

C Cl

O

O

P

OCH3

C

Cl

OCH3

Cl H H

H

tetrachlorvinphos GSH

Cl Cl

O C

Cl

O

Cl

P OCH3

C

Cl

OH

H

Cl

CCH2 Cl

S

G

O

Figure 65.13  Metabolism of dichlorvos and tetrachlorvinphos. GSH, reduced glutathione.

and others are not, as well as the fact that the efficiencies of bioactivations and of detoxications vary substantially among compounds, impart to the organophosphorus insecticides a very wide range of mammalian acute toxicity levels.

References Aizawa, H. (1982). “Metabolic Maps of Pesticides,” p. 243. Academic Press, New York. Aizawa, H. (1989). “Metabolic Maps of Pesticides, Vol. 2.” p. 272. Academic Press, New York. Aldridge, W. N. (1953). Serum esterases: Two types of esterase (A and B) hydrolyzing p-nitrophenyl acetate, propionate and butyrate, and a method for their determination. Biochem. J., 53, 110–117. Appleton, H. T., and Nakatsugawa, T. (1972). Paraoxon deethylation in the metabolism of parathion. Pestic. Biochem. Physiol., 2, 286–294. Chambers, H. W. (1992). Organophosphorus compounds: An overview. In “Organophosphates: Chemistry, Fate and Effects,” (J. E. Chambers and P. E. Levi, Eds.), pp. 3–18. Academic Press, San Diego. Chambers, J. E. (2008). PON1 multitasks to protect health. Proc. Natl. Acad. Sci. USA, 35, 12639–12640. Chambers, H. W., Brown, B., and Chambers, J. E. (1990). Non-catalytic detoxication of six organophosphorous compounds by rat liver homogenates. Pestic. Biochem. Physiol., 36, 308–315.

Chambers, J. E., and Levi, P. E. (ed.) (1992). “Organophosphates: Chemistry, Fate and Effects.” Academic Press, San Diego, California, p. 443. Chambers, J. E., Ma, T., Boone, J. S., and Chambers, H. W. (1994). Role of detoxication pathways in acute toxicity levels of phosphorothion­ ate insecticides in the rat. Life Sci., 54, 1357–1364. Dikshith, T. S. S. Ed. (1991). “Toxicology of Pesticides in Animals.” CRC Press, Boca Raton, Florida. Ecobichon, D. J. (2001). Toxic effects of pesticides. In “Casarett and Doull’s Toxicology: The Basic Science of Poisons,” (C. D. Klaassen Ed.) 6th ed., pp. 763–810. McGraw-Hill, New York. Forsyth, C. S., and Chambers, J. E. (1989). Activation and degradation of the phosphorothionate insecticides parathion and EPN by rat brain. Biochem. Pharmacol., 38, 1597–1603. Levi, P. E., and Hodgson, E. (1992). Metabolism of organophosphorus com­ pounds by the flavin-containing monooxygenase. In “Organophosphates: Chemistry, Fate and Effects,” (J. E. Chambers and P. E. Levi, Eds.), pp. 141–154. Academic Press, San Diego. Levi, P. E., Hollingworth, R. M., and Hodgson, E. (1988). Differences in oxidative dearylation and desulfuration of fenitrothion by cytochrome P450 isozymes and in the subsequent inhibition of monooxygenase activity. Pestic. Biochem. Pharmacol., 32, 224–231. Meister, R. T. Ed. (2004). “Farm Chemicals Handbook 1990,” Meister Publishing Company, Willoghby, Ohio, pp. C1–C496. Neal, R. A. (1980). Microsomal metabolism of thiono-sulfur com­ pounds: mechanisms and toxicological significance. In “Reviews in

Chapter | 65  The Metabolism of Organophosphorus Insecticides

Biochemical Toxicology” (E. Hodgson, J. R. Bend, and R. M. Philpot, Eds.)Vol. 2, pp. 131–172. Elsevier/North Holland, New York. Parkinson, A. (2001). Biotransformation of xenobiotics. In “Casarett and Doull’s Toxicology: The Basic Science of Poisons,” (C. D. Klaassen Ed.) 6th ed., pp. 133–224. McGraw-Hill, New York. Pond, A. L., Chambers, H. W., Coyne, C. P., and Chambers, J. E. (1998). Purification of two rat hepatic proteins with A-esterase activity toward chlorpyrifos-oxon and paraoxon. J. Pharmacol. Exp. Ther., 286, 1404–1411. Pond, A. L., Coyne, C. P., Chambers, H. W., and Chambers, J. E. (1998). Identification and isolation of two rat serum proteins with

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A-esterase activity toward paraoxon and chlorpyrifos-oxon. Biochem. Pharmacol., 52, 363–369. Sultatos, L. G. (1992). Role of glutathione in the mammalian detoxication of organophosphorus insecticides. In “Organophosphates: Chemistry, Fate and Effects,” (J. E. Chambers and P. E. Levi, Eds.), pp. 155–168. Academic Press, San Diego. Tang, J., and Chambers, J. E. (1999). Detoxication of paraoxon by rat liver homogenate and serum carboxylesterases and A-esterasers. J. Biochem. Mol. Toxicol., 13, 261–268.

Chapter 66

Organophosphorus Insecticide Pharmacokinetics Charles Timchalk Pacific Northwest National Laboratory, Richland, Washington

66.1  Background In this chapter an overview will be presented of the pharmacokinetic principles that are of major importance in understanding the toxicology of organophosphorus (OP) insecticides in animals and humans. The approach will not entail a comprehensive review of the extensive literature, but rather a focused presentation highlighting important principles by utilizing specific examples for this class of insecticide. Organophosphates constitute a large family of insecticides that are structurally related pentavalent phosphorus acid esters. Their insecticidal as well as toxicological mode of action is primarily associated with their ability to target and inhibit the enzyme acetylcholinesterase (AChE) (Sultatos, 1994). In this regard, the acute toxic effects of organophosphorus insecticides are associated with the capacity of the parent chemical or an active metabolite to inhibit AChE enzyme activity within nerve tissue (Murphy, 1986; Sultatos, 1994). The three major classes of organophosphorus insecticides are the phosphorothionates, phosphorodithioates, and phosphoroamidothiolates (Chambers, 1992; Mileson et al., 1998). As an example, phosphorothionate insecticides such as chlorpyrifos, parathion, and diazinon are weak inhibitors of AChE, but once they undergo metabolic activation (desulfation) to their corresponding oxygen analogues (oxon) they become extremely potent. This enhanced toxicity is due to the oxon having a high affinity and potency for phosphorylating the serine hydroxyl group within the active site of AChE (Mileson et al., 1998; Sultatos, 1994). The toxic potency is dependent upon the balance between a delivered dose to the target site and the rates of bioactivation and/or detoxification (Calabrese, 1991). The pharmacokinetics and biochemical interactions between organophosphates and AChE and the toxicological implications of AChE inhibition are well understood. To further illustrate this point, a diagram relating insecticide toxicity with pharmacokinetic disposition and the formation of key metabolites is presented in Figures 66.1 and 66.2. The thionophosphate pesticide diazinon [O,O-diethyl-O (2-isopropyl-4-methyl-6-pyrimidinyl) Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

phosphorothioate] is being utilized for illustration purposes; however, based on a common mode of action this scheme is readily extended to other organophosphates. Organophosphorus insecticides, like most chemical contaminants, are absorbed into the body, and, based on the detection of low levels of metabolites in urine within humans, there is good evidence for widespread although low-level exposures (Hill et al., 1995; Aprea et al., 1999). These exposures can come from numerous sources. For example, ingestion of pesticide residues on foods may account for some of the low-level body burdens detected, whereas accidental or intentional ingestion of organophosphorus insecticides is associated with acute poisoning resulting in significantly higher blood, tissue, and urine concentrations of relevant metabolites (Drevenkar et al., 1993). Dermal represents a potential exposure route during the mixing, loading, and application of insecticides or from skin contact with contaminated surfaces (Knaak et al., 1993). Likewise, inhalation of airborne insecticide is feasible either during an application or as the result of exposures associated with chemical drift (Vale and Scott, 1974). Once the organophosphate arrives at a portal of entry it is available for absorption, and, based on the bioavailability for a given insecticide and exposure route, a systemic dose of the parent compound (Figure 66.2, #1) will enter the systemic circulation. Although localized portal of entry metabolism (i.e., lung, intestines, skin) is feasible, the bulk of the metabolic activation as well as detoxification reactions occurs within the liver (Sultatos et al., 1984; Sultatos, 1988). As previously mentioned, phosphorothionates like diazinon do not directly inhibit AChE, but must first be metabolized to the corresponding oxygen analog (oxon; Figure 66.1, #2) (Iverson et al., 1975; Mücke et al., 1970; Murphy, 1986; Sultatos, 1994). Activation to the oxon-metabolite (#2) is mediated by cytochrome P450 mixed function oxidases (CYP450) primarily within the liver, although extrahepatic metabolism has been reported in other tissues including the brain (Chamber and Chambers, 1989; Guengerich, 1977). In addition, oxidative dearylation of the parent compound forming both 2-isopropyl-4-methyl-6-hydroxypyrimidine 1409

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1410

S C2H5O C2H5O

P O

N

CH(CH3)2

N

1

Parent compound

CH3 Diazinon S

O Oxon metabolite

C2H5O C2H5O

2

P O

N

CH3 Diazinon-oxon C2H5O C2H5O

d B-

este

O

S Nonspecific C2H5O P OH 4 metabolite C2H5O Diethylthiophosphate (DETP) CH(CH3)2

CYP450

CYP450 CH(CH3)2 N A-an

rase

HO

N

N

P OH

CH3 2-isopropyl-4-methyl-6hydroxypyrimidine (IMHP)

Diethylphosphate(DEP) 4 Nonspecific metabolite

Major compound specific metabolites

3

CYP450 CH2-OH HO

N

CH N

CH3

CH3 FIGURE 66.1 Metabolic scheme for the metabolism of the organophosphorus (OP) insecticide diazinon. CYP450, cytochrome P450.

Stomach

Oral

1

1

dose Inhalation dose 1

1

Liver 1 2 3 4 •CYP450 –Activation –Detoxification •A-EST (detoxification) •CaE (detoxification)

Other tissues • Fat • Muscle

1

Lung

Skin

Systemic circulation 1 •A-EST (detoxification) •CaE (detoxification) •Blood AChE inhibition 3

2

3

4

1

2 Target tissue AChE inhibition

4

3

4

Kidney

Urine 3

4

Dermal dose 1 FIGURE 66.2 Compartmental flow diagram illustrating the critical tissue compartments associated with absorption, distribution, metabolism, and excretion of organophosphorus (OP) insecticides. The circled numbers (1–4) correspond to the parent compound and major metabolic products associated with metabolism of diazinon (see Figure 66.1) that are most likely found within each compartment. CYP450, cytochrome P450; A-EST, A-esterase; CaE, carboxylesterase; AChE, acetylcholinesterase.

Chapter | 66  Organophosphorus Insecticide Pharmacokinetics

(IMHP, #3), and diethylthiophosphate (DETP, #4), represents a competing detoxification pathway that is likewise mediated by hepatic CYP450 (Ma and Chambers, 1994). These initial activation/detoxification reactions are believed to share a common phosphooxythiran intermediate and represent critical biotransformation steps required for toxicity (Neal, 1980). Differences in the ratio of activation to detoxification are associated with chemical-, species-, gender-, and age-dependent sensitivities to organophosphorus insecticides (Ma and Chambers, 1994). Hepatic and extrahepatic (i.e., blood and tissue) A-esterase (PON1) can effectively metabolize the oxon-metabolite (#2) forming IMHP (#3) and diethylphosphate (DEP, #4) metabolites. Likewise, B-esterases such as carboxylesterase (CaE) and butyrylcholinesterase (BuChE) that are also well distributed across tissues can metabolize the oxon; however, these B-esterases become irreversibly bound (1:1 ratio) to the oxon and thereby become inactivated (Chandra et al., 1997; Clement, 1984). It is likewise clear from both tissue distribution and partitioning studies that phosphothionate pesticides are generally well distributed in tissue throughout the body (Tomokuni, et al., 1985; Wu et al., 1996). Finally, due to the extensive metabolism little if any parent phosphothionate or oxon is available for excretion; however, more stable metabolites such as DEP, DETP, and IMHP are readily excreted in the urine (Iverson et al., 1975; Mücke et al., 1970). Numerous pharmacokinetic approaches have been applied to organophosphorus insecticides, including: 1. Application of pharmacokinetics to understand the overall disposition and clearance 2. Development and application of pharmacokinetic models for quantitative biological monitoring to assess insecticide exposure in humans 3. Studies that facilitate extrapolation of dosimetry and biological response from animals to humans and the assessment of human health risk To illustrate the utility of pharmacokinetics to address health concerns associated with organophosphorus insecticides, several examples of these types of pharmacokinetic studies with these insecticides will be used to illustrate both their utility as well as their limitations.

66.2  Pharmacokinetic principles of importance to organophosphorus insecticides Pharmacokinetics is concerned with the quantitative integration of those processes associated with the absorption, distribution, metabolism, and excretion (ADME) of drugs and xenobiotics within the body (Renwick, 1994). Studies on the pharmacokinetics of a xenobiotic provide critically useful insights into the toxicological response associated with a

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given agent. In this regard, pharmacokinetics provides quantitative data on the amount of toxicant delivered to a target site as well as species-, age-, and gender-specific as well as dose-dependent differences in biological response. An important application of pharmacokinetics within toxicology has been to provide a realistic estimate of risk by providing a means to quantitatively estimate the absorbed dose of a chemical under realistic exposure conditions (Clewell, 1995). Toxicology studies are designed to provide a quantitative assessment of toxicity based on what the chemical agent does to the test animals. In contrast, pharmacokinetics focuses on what the animal does to the chemical. Clearly, toxicity and pharmacokinetics are integrally related since the extent of absorption, retention, metabolic activation, or detoxification is ultimately responsible for delivering a dose to a target tissue resulting in observed effects. Pharmacokinetics represents a critically important tool that, if used correctly, can quantitatively establish a unifying model that describes both dosimetry and biological response across exposure routes, species, and chemical agents. This approach is particularly useful for organophosphorus insecticides since they share a common mode of action through their capability to inhibit AChE activity (Mileson et al., 1998). Pharmacokinetic strategies for quantitating dosimetry can be developed to measure the parent compound and active (i.e., oxon) or inactive metabolites. It is also feasible to link dosimetry with biologically based pharmacodynamic (PD) response models based on a common mode of action (i.e., AChE inhibition). In general, pharmacokinetic modeling approaches can be characterized as empirical or physiologically based, and both types of models have been applied to understand the toxicological response to organophosphorus chemicals in multiple species (Brimer et al., 1994; Gearhart et al., 1990; Pena-Egido, 1988; Poet et al., 2004; Sultatos, 1990; Timchalk et al., 2002a,b, 2005, 2006, 2007a,b; Timchalk and Poet, 2008; Tomokuni et al., 1985; Wu et al., 1996).

66.2.1  Compartmental Pharmacokinetic Models Compartmental models have formed the cornerstone of pharmacokinetic analysis and as such have been extensively utilized to assess bioavailability, tissue burden, and elimination kinetics in various species including humans. All pharmacokinetics are concerned with the time course by which a chemical is absorbed into the systemic circulation, distributed throughout the body, altered through metabolic transformation, and eliminated. Compartmental models are empirical and as such consider the organism as a single or multicompartment homogenous system. The number and behavior of the compartments are primarily determined by the equations chosen to describe the time course data and not the physiological characteristics of the organism (Krishnan and Andersen, 1994). In these models the net transfer between compartments is directly proportional

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to the difference in chemical concentration between compartments. However, the rate constants associated with the transfer between compartments cannot be experimentally determined (Srinivasan et al., 1994). Compartmental models range from a simple well-mixed single compartment to more complicated multicompartment models that are used to describe the blood and/or plasma time course of a chemical or drug. These simple compartmental approaches have been broadly utilized to model the pharmacokinetics of organophosphorus insecticides and their major metabolites (Braeckman et al., 1983; Drevenkar et al., 1993; Nolan et al., 1984; Timchalk et al., 2007b; Wu et al., 1996). For example, Nolan et al. (1984) developed a one-compartment pharmacokinetic model that accurately describes the blood and urine time course of 3,5,6-trichloropyridinol (TCP), a major metabolite of the organophosphorus insecticide chlorpyrifos in human volunteers. A diagram of this single compartment model is illustrated in Figure 66.3. In this model the blood TCP concentration and urinary excretion data were simultaneously fit to a single compartment model using equations 1 and 2. Absorption (ka) and elimination (ke) are handled as firstorder processes, and the blood TCP concentration is represented by Cb, while F and Vd represent fractional absorption and the volume of distribution, respectively. To develop this model, male volunteers were orally administered a 0.5 mg

Oral dose (mg/kg)

ka(h–1)

Cb(µg/ml) =

Cb(µg/ml) =

Absorbed dose (µg)

ke(h–1)

Volume distribution Vd(ml)

Ka � dose � F Vd � (ka–ke)

Urine

� exp(ke � time – ka � time) (1)

Urinary excretion rate (µg/h) = Cb � ke � Vd � body weight (2) Figure 66.3  Single compartment model used to describe the blood and urine time course of 3,5,6-trichloropyridinol (TCP), a major metabolite of the organophosphorus (OP) insecticide chlorpyrifos (CPF) (equations adapted from Nolan et al., 1984).

chlorpyrifos/kg of body weight dose then blood and urine specimens were collected at specified intervals and analyzed for TCP. The model parameters used to describe the time course of TCP and the model fit of the experimental data are presented in Table 66.1 and Figure 66.4. The model provides an excellent fit of the experimental data, and based on the model parameters it was determined that 72% of the ingested dose was absorbed and eliminated in the urine with a half-life of 27 h. Based on this model Nolan et al. (1984) has suggested that blood TCP concentration and/or urinary excretion rate could be utilized to quantify the amount of chlorpyrifos absorbed under actual use conditions. Although compartment modeling is extremely useful for interpolation within the confines of the test species and experimental conditions (i.e., exposure routes and dose levels) these models are limited in their capability to extrapo­ late across dose, species, and exposure routes (Krishnan and Andersen, 1994). To enable extrapolation, physiologically-based pharmacokinetic (PBPK) models have emerged as an important tool that has seen broad applications in toxicology and more specifically in human health risk assessment (Andersen, 1995; Clewell and Andersen, 1996; Krishnan and Andersen, 1994; Leung and Paustenbach., 1995; Mason and Wilson, 1999).

66.2.2  Physiologically-Based Pharmacokinetic Models Unlike compartment modeling approaches, PBPK models utilize biologically meaningful compartments that represent individual organs such as liver and kidney or groups of organ systems (i.e., well perfused/poorly perfused) (Mason and Wilson, 1999). The general model structure is based on an understanding of comparative physiology and xenobiotic metabolism, a chemical’s physical properties that define tissue partitioning, the rates of biochemical reactions determined from both in vivo and in vitro experimentation, and the physiological characteristics of the species of interest (Krishnan and Andersen, 1994). PBPK models have been developed to describe target tissue dosimetry for a broad range of environmental contaminants such as solvents, heavy

Table 66.1  Selected Model Parameters Describing Blood Concentrations and Urinary Excretion of 3,5,6-Trichloropyridinol (TCP) by Individual Volunteers following Oral Administration of the Organophosphate (OP) Insecticide Chlorpyrifos Parameter

Body weight (kg)

Absorption lag time (h)

Absorption rate constant ka (h1)

Absorption half-life (h)

Volume distribution (Vd) (ml/kg)

Elimination rate constant (h1)

Elimination half-life ke (h)

Model predicted dose absorbed (%)

Dose recovered in urine (%)

Range

72–102

0.9–1.9

0.1–2.7

0.4–6.9

160–204

0.02–0.03

21–32

52–84

49–81

Mean  S.D.

83.3   10.3

1.3  0.4

1.5  1.2

0.5

181  18

0.026   0.005

26.9

72  11

70  11

Data obtained from six male volunteers (data adapted from Nolan et al., 1980).

Chapter | 66  Organophosphorus Insecticide Pharmacokinetics

metals, and pesticides, including organophosphorus insecticides (Andersen et al., 1987; Corley et al., 1990; Gearhart et al., 1990; O’Flaherty, 1995; Sultatos, 1990; Timchalk et al., 2002a). A number of reviews have been published on the development, validation, application, and limitations of PBPK models in human health risk assessment (Andersen, 1995; Clewell, 1995; Clewell and Andersen, 1996; Frederick, 1995; Krishnan and Andersen, 1994; Leung and Paustenbach, 1995; Mason and Wilson, 1999; Slob et al., 1997). With regards to the application of PBPK modeling to organophosphorus insecticides, Gearhart et al. (1990) developed a basic PBPK/PD model structure that described target tissue dosimetry and AChE inhibition following an acute exposure to diisopropylfluorophosphate in mice and rats. In developing this model the authors were primarily interested in building a structure that could readily be extended to describe the acute effects for a broad range of commercially important organophosphorus insecticides. A diagram of the PBPK and PD model for diisopropylfluorophosphate in rats is illustrated in Figures 66.5 and 66.6. The conceptual representation of the PBPK model for diisopropylfluorophosphate is based on the anatomical and physiological characteristics of the rat and the major determinants of diisopropylfluorophosphate disposition, which include esterase binding and hydrolysis, tissue partitioning, and diisopropylfluorophosphate volatility (Gearhart et al., 1990; Krishnan and Andersen, 1994). Since this organophosphorus-ester does not require metabolic activation, like thionophosphate insecticides, the hydrolysis of diisopropylfluorophosphate by blood and tissue A-esterase (PON1) is a major factor in determining the protection against AChE inhibition. Diisopropylfluorophosphate binds to and inhibits Besterases including AChE, BuChE, and CaE. Although

1413

binding to AChE is associated with acute neurotoxicity, the binding to BuChE and CaE is without adverse physiological effect and as such represents a detoxification pathway (Clement, 1984; Fonnum et al., 1985; Pond et al., 1995). The PBPK/PD model compartments included those tissues associated with toxicological response (i.e., brain, lung, diaphragm), those containing high A-esterase (PON1) activity (i.e., liver, kidney, and blood), and a fat compartment having the highest tissue/blood partitioning, and the remaining tissues were collectively lumped (Gearhart et al., 1990). To develop this model, tissue partitioning coefficients (PCs) were determined by the vial equilibration technique (Gargas et al., 1989; Sato and Nakajima, 1979). The generalized mass balance differential equations for calculating diisopropylfluorophosphate tissue concentration and AChE tissue inhibition are also presented in Figures 66.5 and 66.6. Within each tissue compartment the net concentration of diisopropylfluorophosphate (mg/l) is a function of blood flow to the tissue, chemical partitioning from the blood into the tissue, and the loss of diisopropylfluorophosphate due to hydrolysis by A-esterase (PON1) and inhibition of B-esterases (AChE, BuChE and CaE). Gearhart et al. (1990) calculated basal AChE activity (mol) based on a zero-order enzyme synthesis rate (mol/h) and a first-order rate of enzyme degradation (h1). A balance between the bimolecular rate of inhibition and the rate of AChE regeneration and aging determined the amount of free AChE. Similar equations were utilized to quantify the impact of diisopropylfluorophosphate on tissue CaE and BuChE activity. The PBPK/PD model developed by Gearhart et al. (1990) was also used as a framework for development of a model for the organophosphorus insecticide chlorpyrifos and diazinon (Poet et al., 2004; Timchalk et al., 2002a). The

Figure 66.4  Time course of 3,5,6-trichloropyridinol (TCP) in the blood and urine of male volunteers orally administered 0.5 mg chlorpyrifos (CPF)/kg of body weight (adapted from Nolan et al., 1984).

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1414

iv and s.c.dose

Mass Balance Differential Equations for DFP Dosimetry

Brain

Venous blood

Liver Kidney Rapid perfused Fat Slow perfused Diaphragm

Qc Qbr Ql Qk Qr

V�dC/dt = Q� (CA – CV)

(DFP tissue concentration) (3)

– (Vmax�Cv)/Km + Cv)

(DFP hydrolysis by A-EST (4)

– KAChE � CAE � C

(DFP inhibition of AChE)

(5)

– KCaE � CCE � C

(DFP inhibition of CaE)

(6)

– KBChE � CBE � C

(DFP inhibition of BChE)

(7)

where V C Q CA CV Vmax Km KAChE CAE KCaE CCE KBChE CBE

Arterial blood

Lungs

Qf Qs Qd

= = = = = = = = = = = = =

Volume of tissue (l) DFP concentration in tissue (mg/l) Blood flow to tissue (l/h) DFP concentration in arterial blood entering tissue (mg/l) DFP concentration in venous blood leaving tissue (mg/l) Maximum rate of A-EST hydrolysis of DFP (mg/h) Michaelis constant for A-EST in tissue (mg/l) Bimolecular rate constant for DFP reaction with AChE (µM h)–1 AChE tissue concentration (µM) Bimolecular rate constant for DFP reaction with CaE (µM h)–1 CaE tissue concentration (µM) Bimolecular rate constant for DFP reaction with BChE (µM h)–1 BChE tissue concentration (µM)

Figure 66.5  Physiologically-based pharmacokinetic (PBPK) model structure and mass balance differential equations describing the distribution of diisopropylfluorophosphate (DFP) in the rat (adapted from Gearhart et al., 1990). Synthesis of new AChE Bimolecular rate of AChE inhibition DFP

Inhibited AChE

Free AChE

Rate of AChE aging

“Aged” AChE

Regeneration of bound AChE Differential Equations for AChE Inhibition Basal degradation of AChE

V × dAE/dt = (KAChE× CAE× C)

(Inhibition of AChE)

– (KRA × AE)

(Regeneration of AChE) (9)

– (KAA × AE)

(Aging of AChE)

where V C KAChE CAE KRA AE KAA

= = = = = = =

(8) (10)

Volume of tissue (l) DFP concentration in tissue (mg/l) Bimolecular rate constant for DFP reaction with AChE(µM h)–1 Free AChE tissue concentration (µM) Rate of regeneration of inhibited AChE (h–1) Inhibited AChE (µM) Rate of aging of inhibited AChE (h–1)

Figure 66.6  Pharmacodynamic (PD) model structure and mass balance differential equations describing the inhibition of acetylcholinesterase (AChE) by diisopropylfluorophosphate (DFP) in the rat (adapted from Gearhart et al., 1990).

key metabolites of chlorpyrifos are illustrated in Figure 66.7. Chlorpyrifos is a phosphorothionate insecticide like diazinon; therefore, they have similar metabolic activation and detoxification reactions (see Figure 66.1). Specifically,

chlorpyrifos undergoes metabolic desulfuration (CYP450) to form the neurotoxic metabolite chlorpyrifos-oxon or dearylation to form 3,5,6-trichloro-2-pyridinol (TCP). A diagram of the PBPK/PD model structure for chlorpyrifos

Chapter | 66  Organophosphorus Insecticide Pharmacokinetics

1415

Cl

AChE Inhibition Toxicity

O

P O

s 450

O O

Chlorpyrifos-oxon

P O

N

A-e

ste ra B-e se (P ste O ras N-1) e

Cl

CH3

Diethylthiophosphate S

CH33 O

P

OH

O

3,5,6-trichloro-2-pyridinol (TCP)

CH3 C

Cl

Cl

CH3

Cl

Chlorpyrifos

Cl

Cl

O

N

CYP450

CH3

O

CH3

P

CY

Cl

S

CH3

O O

Diethylphosphate

P O

OH

OH

CH3

N

Cl

Conjugates- O

Cl

Cl

N

Cl

Sulfate or glucuronides of TCP Figure 66.7  Metabolic scheme for the organophosphorus insecticide chlorpyrifos (CPF).

and chlorpyrifos-oxon is illustrated in Figure 66.8. The major differences between the DFP and chlorpyrifos models is that for chlorpyrifos the model included a PBPK and PD model code for chlorpyrifos-oxon and a compartment model to account for TCP pharmacokinetics. Likewise, metabolic parameters, partition coefficients, and inhibition constants were specifically determined for chlorpyrifos and chlorpyrifos-oxon and used in the model to simulate the pharmacokinetic and pharmacodynamic response in both rats and humans. The capability of the model to simulate both chlorpyrifos and chlorpyrifos-oxon tissue dosimetry and cholinesterase inhibition is illustrated in Figures 66.9 and 66.10, in rats that were administered a range of single oral doses of chlorpyrifos (0.5–100 mg/kg of body weight). The development and application of PBPK modeling for human health risk assessment are not without their challenges and limitations. Before a model can be used to assess risk a determination must be made concerning the model’s capability to accurately predict dosimetry and biological response (Frederick, 1995). Secondly, PBPK/PD models are data intensive, so to adequately develop and validate a model generally requires extensive experimentation to support model parameterization and validation (Clewell, 1995). Nonetheless, a consensus opinion of a panel of expert scientist concluded that biologically based risk assessments that include well-validated PBPK/PD models can provide the most accurate quantitative assessment of human health risk from exposure to environmental chemicals (Frederick, 1995).

66.3  Pharmacokinetic approaches applied to organophosphorus insecticides 66.3.1  Application of Pharmacokinetics to Understand the Overall Disposition and Clearance of Organophosphorus Insecticides Pharmacokinetic studies conducted in multiple species at various dose levels and across different routes of exposure can provide important insight into the in vivo behavior of a chemical agent and how it contributes to the observed toxicological response in a given species. To illustrate this point, a comparison is made of selected pharmacokinetic parameters obtained from a diverse group of studies conducted in animals exposed to either parathion or diazinon. As is noted in Tables 66.2 and 66.3, no single study provides all the pertinent information; yet collectively they provide a consistent qualitative picture of the overall pharmacokinetics of these insecticides. The bioavailability of organophosphorus insecticides, defined as the amount of systemically available dose, is a function of the extent of absorption and first-pass metabolism. Braeckman et al. (1983) conducted a pharmacokinetic study in the dog following both oral and iv administration of parathion. Comparisons of plasma parathion area under the curve (AUC) indicated that 1–29% of the orally administered parathion was bioavailable. The authors suggest that the low systemic oral bioavailability of parathion is primarily

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1416

Skin

Qf

Ca

Chlorpyrifos-oxon Cvsk Cvf

Fat

Qs

Cv

Cvs

Qr

Cvr Rapidly perfused

Qd

Cvd Diaphragm

Qb

Cvb

Brain

Ql

Cvso

Qro Rapidly perfused Qdo

Km1,Vmax1

Dietary Exposure

Hydrolysis Km2,Vmax2

Ke

Cvlo

B-EST (AChE, BuChE, CaE) *Liver and blood only

TCP

Fa

B-Esterase (B-EST) Inhibition (shaded compartments)

Synthesis of New Esterase µmol h–1

“Free” Oxon

Compartment Model

Kzero

Fa

Cvbo

Brain

A-EST* Km3,4, Vmax3,4

Oxon

Intestine

Gavage Exposure

Cvdo

Diaphragm Qbo

Cvro

Liver

KaI

Stomach

Cvo

Slowly perfused

Qlo

Cvl

KaS

Cvfo

Fat

Liver

KsI

Qc

Blood Qfo Qso

Slowly perfused

Arterial Blood

Cao

Venous Blood

Qsk

SA

Arterial Blood

Kp

Venous Blood

Dermal Exposure

Chlorpyrifos Qc

+

Inhibition uM–1h–1

Free Esterase

Oxon-Esterase

h–1

Aged Complex

–1

h Regeneration

h–1

Degradation of Esterase

TCP

Figure 66.8  Physiologically-based pharmacokinetic and pharmacodynamic model used to describe the disposition of the parent insecticide chlorpyrifos (CPF), its oxon metabolite (CPF-oxon), trichloropyridinol (TCP), and B-est inhibition in rats and humans following oral (gavage, dietary) and dermal exposures. The shaded tissue compartments indicate organs in which B-est (AChE, BuChE, and CaE) enzyme activity is described. Model parameter definitions: Qc, cardiac output (l/h); Qi, blood flow to “i” tissue (l/h), Ca, arterial blood concentration (mol/l); Cao, arterial blood concentration of oxon (mol/l); Cv, pooled venous blood concentration (mol/l); Cvi, venous blood concentration draining “i” tissue (mol/l); Cvio, venous blood concentration of oxon draining “i” tissue (mol/l); SA, surface area of skin exposed (cm2); KP, skin permeability coefficient (cm/h); Kzero, zero order (mol/h) rate of absorption from diet; Fa, fractional absorption (%); KaS and KaI, first-order rate constants for absorption from compartments 1 and 2 (per h); KsI, first-order rate constant for transfer from compartments 1 and 2 (per h); Ke, first-order rate constant for elimination of metabolite from compartment 3; Km(1–4), Michaelis constant for saturable processes (mol/l); Vmax(1–4), maximum velocity for saturable process (mol/h).

Chlorpyrifos

B

CPF (µmol/l)

1.E+01 50 mg/kg

1.E+00

10 mg/kg

1.E-01 1.E-02

1 mg/kg

1.E-03 1.E-04 0

10

20

30

Time (h)

CPF-oxon (µmol/l)

A

Chlorpyrifos-oxon 50 mg/kg

1.E-02

10 mg/kg

1.E-03 1.E-04 1.E-05

1 mg/kg

1.E-06 1.E-07

0

10

20

30

Time (h)

Figure 66.9  PBPK/PD model simulations of (A) chlorpyrifos (CPF) and (B) CPF-oxon blood concentrations in adult rats following single oral gavage doses of 1, 10, or 50 mg CPF/kg of body weight. The lines represent the model simulations utilizing the age-dependent PBPK/PD model (experimental data were obtained and adapted from Timchalk et al., 2002 and 2007, respectively, with permission).

associated with a rapid hepatic first-pass metabolism based on the high hepatic extraction (82–97%) that was determined after iv administration. Wu et al. (1996) conducted similar bioavailability studies in the rat with diazinon. The

blood time course of diazinon in the rat following iv and oral doses of 10 and 80 mg/kg, respectively, is presented in Figure 66.11. The results suggest that following oral administration absorption is rapid (absorption t1/2  2.6 h), with

Chapter | 66  Organophosphorus Insecticide Pharmacokinetics

Figure 66.10  Experimental data (symbols) and simulations (lines) from the inhibition of plasma cholinesterase in rats administered chlorpyrifos (CPF) by oral gavage at dose levels of 0.5 (open circle), 1 (open square), 5 (open diamond), 10 (filled circle), 50 (filled triangle), and 100 mg/kg (filled diamond). The data represent the mean  SD of 4 animals per treatment (from Timchalk et al., 2002, with permission).

peak plasma concentrations of diazinon being attained within 2 h postdosing, yet a comparison of AUCs, when corrected for administered dose, indicates that only 35% of the oral dose was bioavailable. The hepatic extraction ratio for diazinon ranged from 48 to 55% and was qualitatively consistent with the findings of Braeckman et al. (1983) for parathion in the dog as well as chlorpyrifos in mice (hepatic extraction ratio 46%) (Sultatos, 1988). A rapid oral absorption (t1/2  0.02 h) and lower oral bioavailability (68%) were also demonstrated in a study where rabbits were administered iv and oral doses of parathion (Pena-Egido et al., 1988). Likewise, in vivo animal models also suggest that dermal absorption and systemic bioavailability of organophosphorus insecticides will be quite low (Brimer et al., 1994). Once these pesticides have been absorbed, systemic distribution throughout the body tissues is rapid (Vale, 1998). For example, a high volume of distribution was observed ranging from 3–14 and 20–23 l/kg, in several different species administered parathion or diazinon, respectively (see Table 66.2). A cross-species comparison of the tissue distribution data following parathion or diazinon exposure is consistent with the large volume of distribution and suggests that the pesticide tissue concentration follows the order of kidney  liver  lung/muscle/heart  brain (see Table 66.3). Phosphorothioates like diazinon and parathion are more lipophilic than their respective oxon metabolites and therefore can be sequestered in the fat compartment, which may account for prolonged intoxication and observed clinical relapses (Vale, 1998). Gearhart et al. (1994) determined the partitioning coefficients (PCs) for both parathion and the toxic metabolite paraoxon (see Table 66.3). In general, the PCs for parathion and paraoxon are comparable; however, parathion has an order of magnitude (101 vs. 10.11) greater affinity than paraoxon for fat. The systemic distribution, elimination kinetics, metabolic transformation, and target site availability of a drug or chemical are often dependent on the extent of reversible plasma/serum protein binding (Renwick, 1994). For example, as is seen in Table 66.2, parathion and diazinon

1417

are extensively bound to plasma protein (ranging from 89 to 99%) and the extent of binding is concentration independent. This response is likewise consistent with the findings of Sultatos et al. (1984), who reported that chlorpyrifos is 97% bound to mouse plasma proteins over a broad concentration range. This high degree of protein binding in conjunction with the high volume of distribution also suggests that tissue binding may in fact be more important that plasma binding in determining the overall disposition and clearance of organophosphorus insecticides (Braeckman et al., 1983). Although the insecticides parathion and diazinon are well distributed throughout the body and extensively bind to both plasma and tissue proteins, they are both rapidly cleared from the body primarily in the urine as degradation metabolites of the parent compounds [i.e., p-nitrophenol, 2isopropyl-4-methyl-6-hydroxyprimidine (IMHP)] (Iverson et al., 1975; Mücke et al., 1970; Nielsen et al., 1991; Vale, 1998). The overall systemic clearance for both parathion and diazinon is quite fast and comparable, ranging from 4 to 6.6 l/h/kg, and is consistent with the rapid blood/plasma terminal phase half-life (2.5–5 h) (see Table 66.2). As previously indicated, comparative species pharmacokinetic analysis is useful for understanding the in vivo behavior of insecticides. Although generalization to all organophosphorus agents is unwise, these types of comparative analyses do provide important insights. In summary, the oral absorption of both parathion and diazinon is rapid, with peak plasma concentrations being obtained within a few hours of exposure. However, oral bioavailability is low and appears to be at least partially associated with a high rate of hepatic first-pass metabolism. Although these insecticides are extensively bound to plasma proteins, they are equally well distributed throughout the body’s tissues, and the parent phosphothioates can sequester within the fat compartment. Nonetheless, the overall clearance is quite fast and is most likely associated with the rapid metabolism and elimination of the metabolites.

66.3.2  Development of Pharmacokinetic Models for Quantitative Biological Monitoring to Assess Organophosphorus Insecticide Exposure in Humans In assessing human exposure to chemical agents, biological monitoring (biomonitoring) is an important quantitative measure of the amount of chemical agent that is systemically absorbed. The approach entails the quantitation of the chemical or its metabolites in biological fluids (i.e., blood, urine, exhaled breath) and offers the best means of accurately assessing exposure since it measures actual, rather than potential, exposure (Woollen, 1993). However, to accurately predict human dosimetry from occupational and/or environmental exposure to xenobiotics, human volunteer

Table 66.2  Selected Model Parameters Describing Blood Concentration Pharmacokinetics of Parent Compounds in Various Species Following Exposure to the Organophosphate (OP) Insecticides Parathion and Diazinon Absorption/bioavailability kinetics

Elimination kinetics

Distribution kinetics

Two-compartment model Species

Dose (mg/kg) Route

Bioavailability (%)

Absorption t1/2 (h)

Hepatic extraction (%)

Volume distribution Vdss (l/kg)

Protein binding (%)

t1/2  (h)

t1/2  (h)

Elimination ke t1/2 (h)

Clearance Cl (l/h/kg)

Rabbitb

1.5 mg/kg iv

100

N/A

—

14.24  6.34

—

—

5.08  3.06

—

3.99  1.13

Rabbit

3 mg/kg oral

68

a

.021  0.04

—

7.58  6.45

—

0.13  0.29

1.08  0.27

2.54  1.67

6.59  3.36

c

0.5 mg/kg iv

100

N/A

—

2.6  0.9

97  1

—

—

3.0  1.5

—

Pig

1 mg/kg iv

100

N/A

—

9.76  5.65

—

—

—

3.6  1.08

4.42  1.20

Pigd

50 mg/kg dermal

9.93  5.28

—

—

—

—

—

—

—

Doge

5 mg/kg iv

—

N/A

82–97

—

99

—

—

—

—

Doge

10 mg/kg oral

1–29

—

—

—

99

—

—

—

—

Ratf

5–10 mg/kg iv

100

N/A

48–55

20.01  11.27

89.1

0.33  0.10

4.70  1.84

—

4.69  0.8

Ratf

80 mg/kg oral

35.5

2.55

—

22.93  4.82

89.1

—

—

2.86  0.58

b

Piglet d

(1) Estimated by comparing oral and iv AUC after adjusting for dose. (2) Data were extracted from the following sources: bPena-Egido et al., 1988; cNielsen et al., 1991; dBrimer et al., 1994; eIverson et al., 1975; fWu et al., 1996. (3) N/A, not applicable.

4.60  1.05

Chapter | 66  Organophosphorus Insecticide Pharmacokinetics

1419

Table 66.3  Tissue Concentration, Tissue Plasma Ratio, and Partition (PC) Coefficients Following Exposure to the Organophosphate (OP) Insecticides Parathion and Diazinon 14

C-Parathiona 0.5 mg/kg; iv Parathionb partition Piglet 3 h postdosing coefficient (PC)

Paraoxonb partition coefficient (PC)

Diazinonc 10 mg/kg; iv Rat 4 h postdosing

Tissues

ng/g

Tissue/ plasma ratio

Tissue/blood Tissue/blood ng/g

Blood/ plasma

262  145

—

—

—

Liver

1254  638

4.78

5.21

Kidney

1360  546

5.19

Lung

421  92

Diazinond 20 mg/kg; iv Mouse 5 h postdosing

Tissue/ plasma ratio

ng/g

Tissue/ plasma ratio

130

—

35

—

6.62

325  25

2.50

120

3.42

5.21

6.62

790  60

6.08

3000

85.7

1.60

5.21e

6.62e

—

—

—

—

f

f

—

—

—

—

Muscle

484  92

1.85

2.55

3.62

Heart

302  85

1.15

—

—

—

—

—

—

Fat

—

—

101.2

10.22

—

—

—

—

Brain

215  76

0.82

4.56

2.31

280  10

2.15

160

4.57

a

Nielsen et al., 1991. Gearhart et al., 1994. c Wu et al., 1996. d Tomokuni et al., 1985. e Well perfused tissue. f Poorly perfused tissue. b

Figure 66.11  Plasma time course of diazinon (DZN) in rats following intravenous (iv) and oral administration of 10 and 80 mg DZN/kg of body weight, respectively (data extracted from Wu et al., 1996).

pharmacokinetic studies conducted under controlled conditions are of vital importance (Wilks and Woollen, 1994; Woollen, 1993). Both occupational and environmental exposure to organphosphorus insecticides is primarily associated with dermal exposure; accounting for more than 90% of the

absorbed dose (Aprea et al., 1994). Therefore, an understanding of the percutaneous absorption is critical for quantitatively determining a systemic dose. The extent of dermal bioavailability for a number of 14C-labeled OP insecticides has been determined in humans utilizing both in vivo studies in volunteers and in vitro dermal penetration with skin

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1420

#2

16

Percent Absorption

14 12 10 8

#3

#1

6

#5 #4

4 2 0

Diazinon (in vivo): Wester et al., 1993 (#1) Diazinon ( in vitro): Wester et al., 1993 (#2) Isofenphos (in vivo): Wester et al., 1992 (#3) Isofenphos (in vitro): Wester et al., 1992 (#4) Malathion (in vivo): Wester et al., 1983 (#5)

Treatment Groups Mean ± S. D. 3.85 ± 2.16 14.1 ± 9.2 3.64 ± 0.48 2.5 ± 2.0 3.53 ± 1.0

Figure 66.12  Summary of human dermal penetration (in vivo/in vitro) for the organophosphate (OP) insecticides diazinon, isofenphos, and malathion.

obtained from cadavers (Wester et al., 1983a, 1992, 1993). A summary of the percent absorption following in vivo and in vitro dermal exposure to the insecticides diazinon, isofenphos, and malathion is illustrated in Figure 66.12. The general experimental design of these studies enta­iled three major components. First, human volunteers were administered a topical dose of a known concentration of 14Clabeled insecticide for a specified exposure period. The extent of absorption was determined by quantitating the amount of 14 C excreted in the urine and remaining on the skin surface. Secondly, in vitro percutaneous absorption was determined using a glass flow-through penetration cell in which the percent absorption though human cadaver skin was determined by the amount of radiotracer that transferred into the receptor fluid. Finally, to calculate the in vivo percent absorption, rhesus monkeys were given a 14C-labeled pesticide as an iv dose. The percent dose absorbed in humans was calculated from the ratio of 14C excreted in the urine after topical (humans) and iv (monkey) dosing. The in vivo absorption for the three insecticides, diazinon, isofenphos, and malathion, in human volunteers following a topical application is very low, ranging from 2.5 to 3.9% of the applied dose. The percent absorption as determined in vitro was likewise comparable for isofenphos (3.64%  0.48), but slightly higher and considerably more variable for diazinon (14.1%  9.2). Percutaneous absorption studies conducted in humans are of particular importance since it is known that dermal absorption in animals, such as the rat, is often greater than in humans (Wester and Maibach, 1983b). For example, Knaak et al. (1990) conducted a dermal absorption study in rats with isofenphos and reported that 47% of the applied dose was absorbed, which is 12-fold higher that the results seen in human volunteers. The major limitation associated with the experimental design of Wester et al. (1983a, 1992, 1993) is

that the quantitation of only 14C provides no information on the specific form of the compound (i.e., parent or metabolite) that is systemically available. Nonetheless, these studies provide important quantitative information on the extent of dermal absorption. To better understand the systemic pharmacokinetics of organophosphorus insecticides and to develop pharmacokinetic models that can be utilized for biomonitoring, controlled human studies that quantitate the time course of parent chemical or metabolites in blood and urine are key. Nolan et al. (1984) conducted a controlled human pharmacokinetic study to follow the fate of a major metabolite, 3,5,6-trichloropyridinol (TCP), which is excreted in the urine following both oral and dermal administration of chlorpyrifos. Griffin et al. (1999) also conducted a controlled human study with chlorpyrifos in human volunteers, but quantitated the urinary excretion of the dialkylphosphate metabolite. A selection of comparative pharmacokinetic parameters from the controlled human chlorpyrifos studies is presented in Table 66.4. Overall, the pharmacokinetic results obtained using TCP or dialkylphosphate in human volunteers are entirely consistent with each other. For example, following oral administration, chlorpyrifos is rapidly absorbed with maximum plasma concentration and excretion being obtained by 6 and 7 h postdosing, respectively, for TCP and dialkylphosphate. The extent of absorption was quite good based on the amount of metabolite (70–93%) recovered in the urine. In comparison, the dermal absorption was consistently slower, with peak concentrations of metabolite being achieved by 17–24 h postdosing for both studies. Also, the amount recovered based on TCP and dialkylphosphate metabolites in the urine was 1.35 and 1%, suggesting limited dermal absorption of chlorpyrifos. Nolan et al. (1984) reported an elimination half-life of

Chapter | 66  Organophosphorus Insecticide Pharmacokinetics

1421

Table 66.4  Comparison of Oral and Dermal Pharmacokinetic Parameters Describing the Blood Concentration and Urinary Excretion of 3,5,6-Trichloropyridinol (TCP) and Dialkylphosphate (DAP) by Volunteers Following Exposure to the Organophosphate Insecticide Chlorpyrifos Exposure route/ metabolite

Dose (mg/kg)

Absorption rate (ng/ cm2/h)

Absorption rate constant ka (h1)

Absorption half-life (h)

Elimination rate constant ke (h1)

Elimination Modelhalf-life (h) predicted dose absorbed (%)

Dose recovered in urine (%)

0.5

—

1.5  1.2

0.5

0.0258  0.0051

26.9

72  11

70  11

0.014

—

—

—

—

15.5

—

93 (range 55–115)

5

—

0.0308  0.01

22.5

—

—

1.35  1.02

1.28  0.83

0.41

456

—

—

—

30

—

1.00

Oral TCPa b

DAP

c

Dermal TCPa DAPb

Data extracted from aNolan et al., 1984; bGriffin et al., 1999. c Estimated based on average body weight (71 kg).

26.9 h following oral administration, whereas Griffin et al. (1999) reported half-lives of 15.5 and 30 h for dialkylphosphate following an oral and dermal exposure to chlorpyrifos, respectively. The increase in the urinary elimination half-life following dermal exposure is most likely associated with a delay in chlorpyrifos absorption through the skin. However, differences in the rates of TCP and dialkylphosphate kinetics are also a possible explanation (Griffin et al., 1999). Nonetheless, the elimination half-life for chlorpyrifos based on either TCP or dialkylphosphate clearance is consistent. These types of pharmacokinetic data are being used to develop models to biomonitor for organophosphate insecticide exposure. Nolan et al. (1984) developed a onecompartment pharmacokinetic model having the same volume of distribution and elimination rate constant to describe blood and urinary TCP kinetics following oral and dermal exposure to chlorpyrifos (see Figure 66.3). Similarly, the quantitative measurement of urinary dialkylphosphate is increasingly being used as a nonspecific biomarker for organophosphorus pesticide exposures (Griffin et al., 1999). Although dialkyl phosphate and TCP have been routinely utilized as biomarkers for insecticide exposure, it is important to acknowledge that organophosphorus pesticides can undergo environmental degradation to form these same chemicals. In this regard, Lu et al. (2005) reported the detection of the breakdown product dialkylphosphate in fruit juices, and Morgan et al. (2005) noted higher concentrations (12–29) of the chlorpyrifos metabolite TCP relative to chlorpyrifos in solid food samples obtained from homes and day care centers; higher dietary exposures to TCP may be a confounding factor when attempting to assess dietary exposure to chlorpyrifos. Hence, due to the environmental stability of the dialkylphosphate and

TCP, recent research has questioned whether total urinary metabolite levels may be reflective of not only an individual’s contact with the parent pesticide, but also exposure with intact metabolites present in the environment (Barr et al., 2004; Bradman et al., 2005; Duggan et al., 2003; Lu et al., 2005). Thus, measured urinary organophosphate metabolite levels may represent an exaggerated indicator of an individual’s exposure to the parent insecticide (Duggan et al., 2003). Nonetheless, the development of pharmacokinetic models that are capable of describing the uptake, distribution, and elimination of insecticides based on the quantitation of major degradation metabolites represents an extremely useful and simple approach for exposure biomonitoring.

66.3.3  The Application of Pharmacokinetics for Quantifying Exposure to Organophosphorus Insecticides The ability to more accurately quantitate human exposure to insecticides has been enhanced by the use of biomonitoring approaches linked to pharmacokinetic analysis. This has successfully been used to estimate agricultural worker exposures during and after the application of insecticides, as an integral component within cross-sectional epidemiology studies to evaluate secondary exposures, and to assess dosimetry in persons who have been acutely poisoned either accidentally or through intentional self-administration (Drevenkar et al., 1993; Lavy et al., 1993; Loewenherz et al., 1997). Historically, workplace exposure to chemicals has been controlled through environmental monitoring that has primarily focused on the measurement of the chemical contaminant

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Total DMTP (µg) in urine

700 600 500 400 300 200 100 0

1

2

3

4

5

6

Amount sprayed (kg a.i.) Figure 66.13  Relationship between the amount of alkyl phosphate (dimethylthiophosphate; DMTP) metabolite in urine of workers and the amount of active ingredient (a.i.) sprayed (data obtained from Franklin et al., 1981).

in the ambient air. However, since airborne concentrations may not be linearly correlated with absorption, this approach does not provide an accurate assessment of internal dose (Franklin et al., 1986). In agricultural settings worker exposure studies have incorporated personal external monitoring to estimate the amount of chemical available from inhalation (i.e., breathing zone sampling pumps) and dermal absorption (i.e., patch method and hand washes). Where feasible, these studies have also incorporated biomonitoring (i.e., urinary metabolites) to quantitate the amount of absorbed dose (Chester, 1993; Franklin et al., 1981, 1986). Franklin et al. (1981, 1986) estimated exposure of workers to the insecticide azinphos-methyl (guthion) utilizing both external personal monitoring and urinary biomonitoring of alkylphosphate metabolites. When patch data were utilized to calculate exposure and plotted against total urinary metabolite excretion, no correlation was observed (Franklin et al., 1981). However, the authors did report a much better correlation when the amount of alkylphosphate metabolite excreted in the urine was compared with the amount of active ingredient sprayed. This relationship is illustrated in Figure 66.13, showing that the amount of alkylphosphate metabolite excreted in the urine increases with increasing amounts of active ingredient. Since agricultural workers routinely apply numerous pesticides and are often sequentially exposed to insecticides within a relatively short time span, a number of exposure studies have been conducted to evaluate mixed insecticide exposures. Hayes et al. (1980) evaluated the occupational exposure of pest control operators in which the bulk of the pesticide applications (80%) involved the combined use of the insecticides vaponite, diazinon, and chlorpyrifos. Worker biomonitoring was based on blood cholinesterase determination and the quantitation of dimethyl- and diethylphosphate and dimethyl- and diethylphosphothioate metabolites in the urine. The authors reported that external air monitoring did provide information regarding the levels and types of exposures, but did not provide adequate

information on the degree to which these insecticides were absorbed. The urinary alkyl phosphate levels provided sensitive quantitative information on absorption and excretion of these pesticides. However, since the alkyl phosphate metabolites are not specific to any one organophosphate insecticide, this approach is indicative only of general exposures to these mixtures and can not be used to quantitatively assess individual insecticide dosimetry. Lavy et al. (1993) conducted a comprehensive yearlong biomonitoring study of tree nursery workers who are routinely exposed to multiple pesticides. In this study it was recognized that as many as 28 pesticides are regularly used, and 17 of the most common pesticides were selected for monitoring, including a number of organophosphates. Evaluation of the human and animal pharmacokinetic data suggested that adequate metabolism information was available on eight of the selected pesticides to support biomonitoring. In this year-long study 3134 urine specimens were analyzed but only 42 of these contained measurable pesticide metabolites (1.3%) and were composed of only three pesticides (benomyl, bifenox, and carbaryl) (Lavy et al., 1993). In addition, based on a calculated margin of safety, the exposure levels were clearly below a level that would be of concern to human health. Biomonitoring strategies have also been successfully applied to quantitatively assess secondary exposures to insecticides resulting from both acute and chronic exposures. Richter et al. (1992) quantitated diethyl phosphate in the urine of individuals who were symptomatic for organophosphate exposure and resided in a house that had been sprayed with diazinon approximately 4.5 months earlier. In this particular study, very high levels of urinary diethyl phosphate were observed in family members, whereas cholinesterase activity, although slightly depressed, was well within the range of “normal.” The quantitation of urinary diethyl phosphate was used to establish a persistent household exposure to diazinon residues as the most likely explanation. This study clearly illustrates the utility of urinary metabolites for quantitative biomonitoring of exposure. Biomonitoring based on the measurement of pesticide metabolites has also been used to compare pesticide exposure in children who live in proximity to high spray areas (e.g., orchards) and whose parents/guardians are pesticide applicators (Loewenherz et al., 1997). Based on known pesticide use patterns, it was determined that insecticide exposure would be primarily associated with azinphos-methyl, chlorpyrifos, and phosmet. Therefore, the study focused on the quantitation of the alkyl phosphate metabolites (dimethylthiophosphate, dimethyldithiophospate, dimethylphosphate) in the children’s urine. Loewenherz et al. (1997) collected and evaluated 160 spot urine specimens from 88 children and reported detectable levels of these metabolites in 27% and 47% of the reference children and applicator children, respectively. In addition, the biomonitoring data suggest that the children of

Chapter | 66  Organophosphorus Insecticide Pharmacokinetics

66.3.4  Studies that Facilitate Extrapolation of Dosimetry and Biological Response from Animals to Humans and the Assessment of Human Health Risk Organophosphorus insecticides constitute a large class of chemical pesticides that are widely used in the agricultural industry and in home applications. This suggests that there is significant potential for exposure, and the health consequences of these exposures may be impacted by both interindividual and extrinsic variability. For example, extrinsic factors such as multiple exposure routes, chemical/drug interactions, and variable exposure rates may significantly modify

µg TCP/l

10

A

1

0.1 0 10 µg chlorpyrifos/l serum

applicators had a significantly higher dose that the reference children (0.021 vs. 0005 g/l, respectively). Biomonitoring of parent pesticide and metabolites in blood and urine has also been used to provide a quantitative assessment of dosimetry in human poisoning victims following acute high-dose exposures (Drevenkar et al., 1993; Vasilic et al., 1992). Although acute cholinesterase depression (i.e., 50% of baseline) is used to substantiate pesticide poisoning, the analysis of intact pesticides or specific metabolites in body fluids (blood/urine) can be used to identify the specific causative chemical agent(s) (Ellenhorn and Barceloux, 1988; Lotti et al., 1986). In this regard, the utilization of pharmacokinetic models like the one developed for chlorpyrifos (Nolan et al., 1984) can be extremely useful for the estimation of dosimetry under these acute exposure scenarios. To illustrate this point, a two-compartment pharmacokinetic model was used to fit pharmacokinetic data obtained from a poison victim who ingested a commercial insecticide formulation containing chlorpyrifos (Drevenkar et al., 1993). These same data have been modeled utilizing a PBPK/PD model developed for the quantitation of chlorpyrifos, chlorpyrifos-oxon, and TCP in the rat and human (Timchalk et al., 2002a). The time course and PBPK/PD model-predicted TCP and chlorpyrifos concentration in the blood and serum of human volunteers and following oral ingestion for a single poison victim is presented in Figure 66.14. The model adequately reflects the data from these limited human samples, but, more importantly, these examples illustrate the strength of using pharmacokinetic models for quantitating dosimetry under both controlled and noncontrolled conditions. In summary, these examples have been presented to illustrate the practical application of pharmacokinetics to assess exposure to chemicals and, more specifically, organophosphorus insecticides. Biomonitoring is clearly an integral component of the agricultural pesticide exposure assessment strategy. However, the successful application of biomonitoring for quantitating dosimetry is primarily limited by a lack of chemical-specific pharmacokinetic data in humans.

1423

10

20

30

40

50 60 Time (h)

70

80

90

100

B

1

0.1

0.01 0

50

100

150

200 250 Time (h)

300

350

400

Figure 66.14  (A) Mean blood time course of 3,5,6-trichloropyridinol (TCP) from six human volunteers administered a single oral dose of 0.5 mg chlorpyrifos (CPF)/kg of body weight (data obtained from Nolan et al., 1984). (B) Time course of CPF in the serum of a single poison victim who orally ingested a commercial insecticide product containing CPF. The symbols represent observed data, while the line represents the model prediction (data obtained from Drevenkar et al., 1993).

the toxicological response to organophosphates. In addition, person-to-person differences in metabolism, genetic predisposition, physical environment, and age (infant, children, and elderly) are important determinants of pharmacokinetic and/or pharmacodynamic response. The development and application of PBPK/PD modeling represent a logical approach to assessing risk and understanding the toxicological implications of known or suspected exposures to various insecticides. The capability of these models to accurately simulate dosimetry and cholinesterase inhibition has been demonstrated in both rodents and humans (Timchalk et al., 2002; Poet et al., 2004). As previously noted, the PBPK/PD model accurately simulates both the pharmacokinetics of chlorpyrifos (Figure 66.9) and the dynamics of cholinesterase inhibition (Figure 66.10) in the rat. Likewise, the model has also been used to simulate human dosimetry and cholinesterase dynamics utilizing pharmacokinetic and pharmacodynamic data obtained from control human studies (Nolan et al., 1984; Timchalk et al., 2002). For example, the time course of plasma BuChE inhibition kinetics

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1424

following a single oral (0.5 mg/kg) or dermal (5 mg/kg) dose of chlorpyrifos in human volunteers is presented in Figure 66.15. These results clearly illustrate the routedependent (oral vs. dermal) differences in the extent of plasma cholinesterase response. The PBPK/PD model also accurately simulated the time course of chlorpyrifos and TCP in blood of human volunteers following oral exposure (1 and 2 mg/kg) to chlorpyrifos (see Figure 66.16). These model simulations are consistent with the rapid metabolic clearance of chlorpyrifos resulting in the formation of TCP, which has a considerably slower elimination rate and is readily detected in the blood of volunteers through 160 h postdosing. With the development and validation of the PBPK/PD model in both rats and humans, the model can now be exploited to quantitatively assess dosimetry and dynamic

response over a range of relevant occupational and environmental chlorpyrifos exposures. Hence, this model is a strong framework for refining a biologically based risk assessment for exposure to chlorpyrifos under a variety of scenarios (Timchalk et al., 2002).

66.3.4.1  Insecticide Mixtures Both occupational and secondary exposures to insecticides often entail simultaneous or sequential contact with mixtures (Hayes et al., 1980; Lavy et al., 1993; Loewenherz, et al., 1997). The potential for organophosphorus insecticide interactions has been well understood for some time. Early studies demonstrated the acute, synergistic, and toxicological interactions between malathion and EPN (ethylp-­nitrophenyl phenylphosphonothionate) (Frawley et al.,

Dermal 5 mg/kg

Figure 66.15  Experimental data (symbols) from Nolan et al. (1984) and model simulations (lines) of the plasma cholinesterase (ChE) inhibition in human volunteers administered an oral dose (filled circles) of 0.5 mg chlorpyrifos (CPF)/kg or a dermal dose (filled squares) of 5 mg CPF/kg. The time course data represent the mean  SD for five male volunteers (adapted from Timchalk et al., 2002, with permission).

Figure 66.16  Experimental data (symbols) and model simulations (lines) for the plasma concentration of trichloropyridinol (TCP) (filled circles) and chlorpyrifos (CPF) (filled squares) in five volunteers administered CPF as an oral dose of 1 mg/kg (A and B) or 2 mg/kg (C and D) (adapted from Timchalk et al., 2002, with permission).

Chapter | 66  Organophosphorus Insecticide Pharmacokinetics

1957). In addition, coexposures to noninsecticides have been reported to influence the pharmacokinetic and toxicological response of organophosphates. For example, phenobarbital or alcohol pretreatment of mice protects against the acute toxicity of chlorpyrifos and parathion, respectively (O’Shaughnessy and Sultatos, 1995; Sultatos et al., 1984). Wu et al. (1996) reported that pretreatment of rats with cimetidine potentiated the acute toxicity of diazinon as a result of reducing diazinon total body clearance. Likewise, coadministration of diazinon with cocaine significantly increased the concentration of cocaine and norcocaine in the blood and tissues of mice, apparently due to competition for esterase enzyme detoxification (Benuck et al., 1989; Kump et al., 1994). A combination of malathion and the carbamate pesticide carbaryl alters the fundamental pharmacokinetic properties of the individual compounds, and it has been suggested that this may explain some of the observed toxicity seen from exposure to this chemical mixture (Waldron Lechner and Abdel-Rahman, 1986). Organophosphorus pesticides as a class of compounds share common metabolic processes for activation and detoxification as well as a common mechanism for toxicological response through the inhibition of AChE (Murphy, 1986; Sultatos, 1994). Based on similar pharmacokinetic and mode of action properties, a potential for interactions between mixtures of these insecticides is hypothesized. Organophosphates can interact at a number of important metabolic steps (see Table 66.5) including: (1) CYP450-mediated activation/ detoxification; (2) plasma protein binding; (3) PON-1 (Aesterase) detoxification; and (4) AChE binding/inhibition. The net effect of these interactions (additivity, synergy, or antagonism) will be dependent upon the specific mixture, dose ranges of exposures, and sensitivity of the individual. To provided needed perspective on organophosphorus insecticide mixture interactions, a binary PBPK/PD model for chlorpyrifos, diazinon, and their metabolites was developed (Timchalk and Poet, 2008) based on previously

1425

published models for the individual insecticides (Poet et al., 2004; Timchalk et al., 2002, 2007a). This model was designed to quantitatively integrate both tissue dosimetry and dynamic response (ChE inhibition) in blood and tissues. In addition, the metabolic interactions (CYP450) between chlorpyrifos and diazinon were evaluated in vitro to characterize the binary mixture enzymatic kinetic interactions for the mixture. Based on the in vitro metabolism results, the PBPK model codes used to describe the CYP450 metabolism of chlorpyrifos and diazinon were appropriately modif­ied to reflect the type of inhibition kinetics (i.e., competitive vs. noncompetitive), while B-est metabolism was described as dose additive, and no PON-1 interactions were assumed between chlorpyrifos- and diazinon-oxon with the enzyme. The binary model was evaluated against previously published rodent dosimetry and cholinesterase inhibition data for the mixture (Timchalk et al., 2005). The metabolic interaction (CYP450) between chlorpyrifos and diazinon was evaluated in vitro and the results indicated that chlorpyrifos was more substantially metabolized to its oxon metabolite than diazinon, which is consistent with the in vivo potency (chlorpyrifos  diazinon). Each insecticide inhibited the other’s in vitro metabolism in a concentration-dependent manner. Based on the differences in oxon formation rates, the most dramatic difference is associated with the extent of cholinesterase inhibition for single insecticides versus. the binary mixtures. To illustrate this point, the time- and mixture-dependent inhibition of brain AChE following single or binary exposures to chlorpyrifos and diazinon is presented in Figure 66.17. The general response for brain AChE and model simulations is also consistent with the pharmacodynamics observed for plasma cholinesterase and RBC AChE inhibition. In the brain there is a potential shift toward shorter times to achieve maximum inhibition with increasing doses of insecticides (Timchalk et al., 2005), which is reasonably simulated by

Table 66.5  Important Metabolic and Response Interactions for Mixtures of Organophosphate (OP) Insecticides Parameters

Importance

Type of Chemical Interaction

Implications

CYP450 mixed-function oxidase metabolism

Metabolic activation/ detoxification of parent compound

Substrate (parent compound) competition for enzyme

Changes in oxon concentrations

Reversible plasmaprotein binding

Systemic transport of parent compound

Substrate (parent compound) competition for available protein binding sites

Increased levels of “free” parent chemical available for metabolism

A-esterase metabolism

Important metabolic step responsible for detoxification

Substrate (oxon) competition for enzyme

Changes in oxon concentrations

AChE binding/inhibition

Toxicological response

Substrates (oxon) combine to increase inhibition of AChE

Increased toxicity due to additive response

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Brain AChE 15 mg/kg 100 80 60 40 20 0

CPF only CPF + DZN

0

30

20 Time (h)

60 mg/kg

10 20 Time (h)

0

10

20

30

30 mg/kg 100 80 60 40 20 0

0

30

10

20

30

Time (h)

F

100 80 60 40 20 0 0

DZN + CPF

30

Percent of control

C

10

DZN only

E

100 80 60 40 20 0 0

100 80 60 40 20 0

Time (h)

Percent of control

Percent of control

20 Time (h)

30 mg/kg

B

Percent of control

10

Brain AChE 15 mg/kg

D Percent of control

Percent of control

A

60 mg/kg 100 80 60 40 20 0

0

10 Time (h)

20

30

Figure 66.17  Brain acetylcholinesterase (AChE) activity in rats following 15, 30, or 60 mg/kg oral gavage doses of chlorpyrifos (CPF) (solid black diamonds), diazinon (DZN) (open black diamonds), and their binary mixtures (solid or open gray circles). The data are expressed as percent of total ChE activity as a function of time (h) and represent the mean  SD for four animals per time point. The lines represent the PBPK/PD model simulations (adapted from Timchalk and Poet, 2008, with permission).

the PBPK/PD model. Consistent with the experimental finding, model simulations suggest that chlorpyrifos has a substantially greater inhibitory impact on brain AChE activity than diazinon at all dose levels. The PBPK/PD model was also utilized to further evaluate theoretical mixture interactions over a very broad range of mixture doses. In these simulations (see Figure 66.18) single acute doses ranging from 1 to 200 mg/kg were evaluated, recognizing that the high end of the dose range (50–200 mg/kg) would result in substantial adverse acute toxicity, including lethality. The rationale for simulating these very high acute doses was to try and establish at what dosage the model would predict a deviation between the single and binary exposures. These simulations are very consistent with the experimental findings and suggest that binary interactions between chlorpyrifos and diazinon at environmentally relevant exposures levels will most likely be negligible and that chlorpyrifos has a greater impact than diazinon as a binary mixture (Timchalk et al., 2008). Based on the results with this PBPK/PD model, it is anticipated that at low binary doses most likely to be encountered in both occupational- and environmental-related

exposures, the pharmacokinetics are expected to be linear, and cholinesterase inhibition kinetics are well described using a dose-additive model. Hence, this binary model provides a mechanistic framework for understanding the lack of important synergistic interactions at occupationally and environmentally realistic exposures, even for pesticides that are as similar as chlorpyrifos and diazinon (Timchalk et al., 2008).

66.3.4.2  Sensitive Subpopulations (Children) There is currently a significant focus on and concern over the potential increased sensitivity of infants and children to the toxic effects of chemicals. The importance of this issue is highlighted by the National Research Council’s report On Pesticides in the Diets of Infants and Children and the Food Quality Protection Act. It is recognized that children are not just “small adults,” but rather a unique subpopulation that may be particularly vulnerable to chemical insult. Agedependent changes in a child’s physiology (i.e., body size, blood flow, organ functions) and metabolic capacity (i.e., phase I and II metabolism) may significantly impact the

Chapter | 66  Organophosphorus Insecticide Pharmacokinetics

D

Blood CPF AUC 300

CPF only

250

CPF + DZN

200 150 100 50 0 0.1

1

10

100

1000

DZN AUC (µmol l−1 h−1)

CPF AUC (µmol l−1 h−1)

A

1427

Blood DZN AUC 70 60 50 40 30 20 10 0

DZN only DZN + CPF

0.1

1

Dose (mg/kg)

E

Blood CPF-oxon AUC 0.16 0.14 0.12 0.1 0.08 0.06 0.04 0.02 0 0.1

1

10

100

1000

Dose (mg/kg)

Brain CPF-oxon AUC

F

1

0.04 0.035 0.03 0.025 0.02 0.015 0.01 0.005 0 0.1

10

100

1000

Blood DZN-oxon AUC 0.014 0.012 0.01 0.008 0.006 0.004 0.002 0 0.1

1

10

100

1000

Dose (mg/kg)

DZN-oxon AUC (µmol l−1 h−1)

CPF-oxon AUC (µmol l−1 h−1)

C

DZN-oxon AUC (µmol l−1 h−1)

CPF-oxon AUC (µmol l−1 h−1)

B

10

Dose (mg/kg)

100

1000

Dose (mg/kg)

Brain DZN-oxon AUC 0.014 0.012 0.01 0.008 0.006 0.004 0.002 0 0.1

1

10

100

1000

Dose (mg/kg)

Figure 66.18  Model simulation of chlorpyrifos (CPF), diazinon (DZN), CPF-oxon, and DZN-oxon area under the concentration curve (AUC) for single and binary exposures to CPF and DZN over a broad range of doses (1–200 mg/kg). The solid gray diamonds represent CPF and DZN as single doses, while the solid black circles represent the binary mixture (adapted from Timchalk and Poet, 2008, with permission).

response to a chemical insult, resulting in either beneficial or detrimental effects (Miller et al., 1997). Clear variability in the capacity to detoxify environmental chemicals has been established in both animals and humans. However, the current risk assessment paradigms do not adequately consider the implications of these differences on the risk to infants and children. Numerous studies have demonstrated that juvenile animals are more susceptible to the acute effects of organophosphorus insecticides than adults (Benke and Murphy, 1975; Brodeur and DuBois, 1963; Gaines and Linder, 1986; Harbison, 1975; Moser and Padilla, 1998; Pope et al., 1991; Pope and Liu, 1997). This greater sensitivity has primarily been attributed to the lack of complete metabolic competence during neonatal and postnatal development (Benke and Murphy, 1975). The application of physiologically based pharmacokinetic/pharmacodynamic (PBPK/PD) modeling offers a unique opportunity to integrate age-dependent changes in metabolic activation and detoxification pathways into a comprehensive model that is capable of quantifying dosimetry

and response across all ages (for review, see Corley et al., 2003). In this context, PBPK models are being extended to the modeling of chemical exposure in developing/juvenile animals and in children (Byczkowski et al., 1994; Clewell et al., 2002, 2003, 2004; Fisher et al., 1990; Sundberg et al., 1998; Price et al., 2003). To address this issue for the organophosphorus insecticides, the PBPK/PD model that was developed for chlorpyrifos in adult rats and humans (Timchalk et al., 2002) was modified by incorporating age-dependent scaling to adjust physiology, organ volumes and blood flows, metabolism rates, B-est tissue levels, and bimolecular inhibition rates for chlorpyrifos-oxon and cholinesterase as a function of age (Timchalk et al., 2007a). The model was then used to predict tissue dosimetry and pharmacodynamic (PD) response (i.e., esterase inhibition) in preweanling and adult rats exposed to chlorpyrifos (Timchalk et al., 2006). To simulate the kinetics of chorpyrifos dosimetry and cholinesterase inhibition in preweanling rats, the PBPK/PD model was modified to scale metabolism, cholinesterase

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II 0.4 0.2 0.1 0

0

10

20

30 40 50 Age (days)

Percent body weight

B 6 5 4 3 2 1 0

Scaling metabolism Liver PON1

A

0.3

60

70

80

(µmol/h)

Body weight (kg)

I A

Liver

100000 10000 1000 100 10 1 0.1 0.01

Plasma PON1 CYP450 dearylation

CYP450 desulfation 0

20

40

60

80

100

Age (days)

0

10

20

30 40 Age (days)

50

60

B

Plasma B-esterase (µmol)

CaE

1.E-01 6 5 4 3 2 1 0

Brain

(µmol)

Percent body weight

C

AChE

1.E-02 1.E 03 1.E-04 1.E-05 1.E-06

BuChE

1.E-07 0 0

10

20

30 40 Age (days)

50

60

10

20

30

40

50

60

Age (days)

Figure 66.19  (I) Age-dependent scaling of (A) body weight (kg), (B) liver volume (percent of body weight), and (C) brain volume (percent of body weight) as a function of age in Sprague-Dawley rats. (II) (A) Scaling of metabolic rates as a function of age for hepatic and plasma PON-1 and hepatic CYP450-mediated dearylation and desulfation of chlorpyrifos (CPF) and CPF-oxon in the rat. Symbols represent experimentally determined enzyme activity (Atterberry et al., 1997), whereas lines represent model prediction. (B) Scaling of the amount of plasma B-est (CaE, AChE, and BuChE) as a function of age in the rat; similar scaling was done for hepatic, diaphragm, and brain B-est levels (from Timchalk et al., 2007, with permission).

activity, and most relevant organ volumes and body weight as a function of age (Timchalk et al., 2007a). Figure 66.19 illustrates the dynamic changes in physical parameters, key metabolism pathways, and cholinesterase levels for the developing preweanling rat. It should be noted that a number of other model parameters were modified to accommodate age-dependent developmental changes; for a more detailed discussion see Timchalk et al. (2007a). Figure 66.20 illustrates the capability of the model to simulate blood concentrations of both chlorpyrifos and TCP in very young preweanling rats following a single bolus or fractioned dose. The overall pharmacokinetic profile was very comparable and the PBPK/PD model reasonably simulated both the chlorpyrifos and TCP blood time course. As anticipated, the Cmax for chlorpyrifos and TCP following the fractionated doses was lower (2.8- to 7.4-fold) than following the single bolus administration, and the PBPK/PD model reasonably simulated this response (Timchalk et al., 2007a). The model was also utilized to simulate the time courses of plasma cholinesterase, RBC AChE, and brain AChE inhibition in adult and preweanling rats (Timchalk

et al., 2002, 2006) following single oral gavage administration of chlorpyrifos (Timchalk et al., 2007a). In the brain, the AChE inhibition demonstrated both an age and dose dependency in the preweanling rats (Timchalk et al., 2006), and at all dose levels and ages the model reasonably simulated the dynamics of brain AChE inhibition (Figure 66.21). Of particular importance was the observation that even in rats as young as postnatal day 5 (PND-5), the CYP450 metabolic capacity was adequate to metabolize chlorpyrifos to both TCP and oxon based on the detection of TCP in blood and extensive cholinesterase inhibition. A comparison of the simulated oxon AUC ratios (PND5 vs. adult) in both blood and brain over a broad range of doses is illustrated in Figure 66.22. At doses ranging from 0.001 to 1 mg/kg the preweanling to adult oxon ratio (PND  5(AUC)/Adult(AUC)) for blood and brain was 1.3; however, at chlorpyrifos doses 1 mg/kg the ratio rapidly increased in both blood and brain and approached 2 at 10 mg/kg. This suggests that age-dependent difference in brain oxon concentration may be an important contributing factor associated with the increased sensitivity of

Chapter | 66  Organophosphorus Insecticide Pharmacokinetics

Cmax (µmol/l) TCP: 1.61 ± 0.56 CPF: 0.140 ± 0.090

1 mg/kg PND-5 1.E+01

TCP

(µmol/l)

1.E+00 1.E-01 1.E-02

CPF

PND-5

A Percent of control

A

1429

MaxI (% Control) 1 mg/kg: 77.9 ±11.8 10 mg/kg: 16.4 ±13.4

0

5

10

1.E-03 10

20

30

Time (h) B

PND-5 split dose 1 mg/kg

Cmax (µmol/l) TCP:0.566 ± 0.193 CPF: 0.019 ± 0.002

1.E+00

1.E-03

C

CPF

1.E-04 1.E-05

0

20

40 Time (h)

60

80

Figure 66.20  Observed data and model prediction (lines) for blood concentrations of chlorpyrifos (CPF) and trichloropyridinol (TCP) in postnatal day 5 (PND-5) rats given (A) a single oral gavage dose of 1 mg/ kg or (B) three split doses of 0.33 mg/kg, each administered at 0, 8, and 16 h. The observed data are a mean  s.d. of four to five animals per time point. The Cmax is the maximally measured blood concentration (adapted from Timchalk et al., 2007, with permission).

preweanling rats relative to adults, particularly at the higher doses utilized in toxicology studies. As these results indicate, an age-dependent PBPK/PD model for the organophosphorus insecticide chlorpyrifos behaves consistently with the general understanding of toxicity, pharmacokinetics, and tissue cholinesterase inhibition in preweanling and adult rats. Future model development must entail further development and validation with the ultimate goal of developing a model that is capable of predicting biological response in infants and children. Nonetheless, the utilization of PBPK/PD modeling to address organophosphorus insecticide toxicity issues is particularly intriguing since these models can be used to assess the health consequences of both interindividual (i.e., age, gender) and extrinsic factors (i.e., multiple exposure routes, chemical/drug interactions, and variable exposure rates) that may significantly modify the toxicological response.

Conclusion This chapter has illustrated a number of current and future applications of pharmacokinetics to assess organophosphorus insecticide dosimetry, biological response, and risk in humans exposed to these insecticides. Pharmacokinetics is concerned with the quantitative integration of absorption, distribution,

30

MaxI (% Control) 1 mg/kg: 94.8 ± 17.5 10 mg/kg: 19.6 ± 5.37

0

1.E-02

25

120 100 80 60 40 20 0

TCP

Percent of control

(µmol/l)

1.E-01

15 20 Time (h) PND-12

B Percent of control

0

5

10

15 20 Time (h)

25

30

PND-17 120 100 80

MaxI (% Control) 1 mg/kg: 97.9 ± 5.23 10 mg/kg: 41.1 ± 4.82

60 40 20 0 0

5

10

15 Time (h)

20

25

30

Adult

D Percent of control

1.E-04

Brain AChE

120 100 80 60 40 20 0

120 100 80 60 40 20 0

MaxI (% Control) 1 mg/kg: 103 ±3.60 10 mg/kg: 84.9 ±3.20

0

5

10

15 Time (h)

20

25

30

Figure 66.21  Observed data and model prediction (lines) for the brain AChE inhibition time course in (A) postnatal day 5 (PND-5), (B) PND12, (C) PND-17, and (D) adult rats following a single acute oral gavage dose of 1 (open diamonds) or 10 (closed triangles) mg CPF/kg of body weight. The observed data are presented as a mean  s.d. of four to five animals per time point. The maximum inhibition (MaxI) is expressed as percent of control activity for each of the dose levels (adapted from Timchalk et al., 2007, with permission).

metabolism, and excretion and can be used to provide useful insight into the toxicological responses associated with these insecticides. Since organophosphorus insecticides share a common mode of action through their capability to inhibit AChE activity, it is feasible to develop pharmacokinetic strategies that link quantitative dosimetry with biologically-based pharmacodynamic (PD) response modeling. Pharmacokinetic studies that have been conducted with organophosphorus insecticides in multiple species, at various dose levels, and across different routes of exposure have provided important insights into the in vivo behavior of these insecticides. The development and application of pharmacokinetic models capable of describing uptake, distribution, metabolism, and elimination of insecticides in humans represent a ­crucial

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ACKNOWLEDGMENTS

A Ratio of blood CPF-oxon AUC (PND-5: Adult)

2

This work was supported partially by Grants R01 OH008173 and R01 OH 003629 from the Centers for Disease Control/ National Institute of Occupational Safety and Health (CDC/ NIOSH). The contents of this publication are solely the responsibility of the authors and do not necessarily represent the official view of CDC/NIOSH. This manuscript has been authored by Battelle Memorial Institute, Pacific Northwest Division, under Contract No. DE-AC05-76RL0 1830 with the U.S. Department of Energy.

1 .8 1 .6 1 .4 1 .2 1 0.0 01

0.01

0.1 1 Dose (mg/kg)

10

100

B Ratio of brain CPF-oxon AUC (PND-5:Adult)

2 1 .8 1 .6 1 .4 1 .2 1 0.0 01

0.01

0.1

1

10

100

Dose (mg/kg) FIGURE 66.22 The ratio of (A) blood and (B) brain chlorpyrifos-oxon (CPF-oxon) area under the concentration curves (AUC0–α) comparing neonatal (PND-5) and adult rats (PND  5(AUC)/Adult(AUC)) following simulation of an acute oral exposure to a broad range of CPF doses (adapted from Timchalk et al., 2007, with permission).

research element needed for quantitative biomonitoring. In this regard, the successful application of biomonitoring for quantitating dosimetry is primarily limited by the lack of this chemical-specific pharmacokinetic data in humans. The development and application of PBPK/PD modeling for organophosphorus insecticides represent a unique opportunity to quantitatively assess human health risk and to understand the toxicological implications of known or suspected exposures to various insecticides. Validated PBPK/PD models for these insecticides can be used to consider the potential variability in human response associated with both interindividual (i.e., age, gender, polymorphism) and extrinsic variability (i.e., exposure routes and rates, single vs. multiple exposures). In conclusion, pharmacokinetics has been successfully utilized to better understand the toxicological implications of human exposure to organophosphorus insecticides. Nonetheless, there is still a significant need to further develop and refine pharmacokinetic models that can be used to accurately assess the risk associated with insecticide exposures.

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Timchalk, C., Busby, A., Campbell, J. A., Needham, L. L., and Barr, D. B. (2007b). Comparative pharmacokinetics of the organophosphorus insecticide chlorpyrifos and its major metabolites diethylphosphate, diethylthiophosphate and 3,5,6-trichloro-2-pyridinol in the rat. Toxicol. 237, 145–157. Timchalk, C., Kousba, A. A., and Poet, T. S. (2007a). An age-dependent physiologically based pharmacokinetic/pharmacodynamic model for the organophosphorus insecticide chlorpyrifos in the preweanling rat. Toxicol. Sci. 98(2), 348–365. Timchalk, C., Kousba, A., and Poet, T. S. (2002b). Monte Carlo analysis of the human chlorpyrifos-oxonase (PON1) polymorphism using a physiologically based pharmacokinetic and pharmacodynamic (PBPK/PD) model. Toxicol. Lett. 135, 51–59. Timchalk, C., Nolan, R. J., Mendrala, A. L., Dittenber, D. A., Brzak, K. A., and Mattsson, J. L. (2002a). A physiologically based pharmacokinetic and pharmacodynamic (PBPK/PD) model for the organophosphate insecticide chlorpyrifos in rats and humans. Toxicol. Sci. 66, 34–53. Timchalk, C., and Poet, T. S. (2008). Development of a physiologically based pharmacokinetic and pharmacodynamic model to determine dosimetry and cholinesterase inhibition for a binary mixture of chlorpyrifos and diazinon in the rat. NeuroTox. 29, 428–443. Timchalk, C., Poet, T. S., Hinman, M. N., Busby, A. L., and Kousba, A. A. (2005). Pharmacokinetic and pharmacodynamic interaction for a binary mixture of chlorpyrifos and diazinon in the rat. Toxicol. Appl. Pharmacol. 205, 31–42. Timchalk, C., Poet, T. S., and Kousba, A. A. (2006). Age-dependent pharmacokinetic and pharmacodynamic response in preweanlng rats following oral exposure to the organophosphorus insecticide chlorpyrifos. Toxicol. 220, 13–25. Tomokuni, K., Hasegawa, T., Hirai, Y., and Koga, N. (1985). The tissue distribution of diazinon and the inhibition of blood cholinesterase activities in rats and mice receiving a single intraperitoneal dose of diazinon. Toxicol. 37, 91–98.

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Vale, J. A. (1998). Toxicokinetic and toxicodynamic aspects of organophospate (OP) insecticide poisoning. Toxicol. Lett. 102-103, 649–652. Vale, J. A., and Scott, G. W. (1974). Organophosphate poisoning. Guy’s Hosp. Gazette 123, 12–25. Vasilic, Z., Drevenkar, V., Rumenjak, V., Stengl, B., and Frobe, Z. (1992). Urinary elimination of diethylphosphorus metabolites in persons poisoned by quinalphos or chlorpyrifos. Arch. Environ. Contam. Toxicol. 22, 351–357. Waldron Lechner, D., and Abdel-Rahman, M. S. (1986). Kinetics of carbaryl and malathion in combination in the rat. J. Toxicol. Environ. Health 18(2), 241–256. Wester, R. C., and Maibach, H. I. (1983b). Cutaneous pharmacokinetics: 10 steps to percutaneous absorption. Drug Metab. Rev. 14, 169–205. Wester, R. C., Maibach, H. I., Bucks, D. A. W., and Guy, R. H. (1983a). Malathion percutaneous absorption after repeated administration to man. Toxicol. Appl. Pharmacol. 68, 116–119. Wester, R. C., Maibach, H. I., Melendres, J., Sedik, L., Knaak, J., and Wang, R. (1992). In vivo and in vitro percutaneous absorption and skin evaporation of isofenphos in man. Fund. Appl. Pharmacol. 19, 521–526. Wester, R. C., Sedik, L., Melendres, J., Logan, F., Maibach, H. I., and Russell, I. (1993). Percutaneous absorption of diazinon in humans. Fd.Chem. Toxic. 31(8), 569–572. Wilks, M. F., and Woollen, B. H. (1994). Human volunteer studies with non-pharmaceutical chemicals: metabolism and pharmacokinetic studies. Hum. Exp. Toxicol. 13(6), 383–392. Woollen, B. H. (1993). Biological monitoring for pesticide absorption. Occup. Hyg. 37(5), 525–540. Wu, H. X., Evreux-Gros, C., and Descottes, J. (1996). Diazinon toxicokinetics, tissue distribution and anticholinesterase activity in the rat. Biomed. and Environ. Sci. 9, 359–369.

Chapter 67

Neuropathy Target Esterase Sanjeeva J. Wijeyesakere and Rudy J. Richardson University of Michigan, Ann Arbor, Michigan

67.1  Introduction Neuropathy target esterase (NTE) is a biochemical enigma. This protein, present in neural and other tissues, has the capacity to catalyze the hydrolysis of certain physiological phospholipids as well as unphysiological carboxylic ester substrates. Like acetylcholinesterase (AChE), its catalytic activity is inhibited by covalent reaction with a variety of progressive inhibitors, including certain organophosphorus (OP) esters. However, unlike AChE, mere loss of the catalytic activity of NTE seems not to be deleterious in adult animals. Rather, the nature of the chemical group covalently bound at the catalytic center appears to determine whether toxicological effects will follow. Forty years after its discovery, there is now overwhelming support for the hypothesis that initiation of OP ester-induced delayed neuropathy/neurotoxicity (OPIDN) by neuropathic OP compounds starts with NTE. Initiation may occur within hours of administration of a single dose of a neuropathic OP ester, although clinical expression is deferred for 1–4 weeks. The clinical and morphological features of OPIDN are described in Chapter 69, as are some of the biochemical and physiological changes reported to accompany the development of the syndrome after initiation. In this chapter, we review the research that led to the identification of NTE as the site of initiation of OPIDN and the possible involvement of related esterases in promotion/ potentiation of the disease. This is followed by a discussion of the application of NTE studies to the assessment of neuropathic risk of OP compounds. We then consider the nature and properties of this remarkable protein, including its possible function in neurons. Finally, we examine two opposing theories concerning the pathogenic role of NTE

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in OPIDN and other neurodegenerative diseases: toxic gain of function versus loss of function with consequent perturbations of lipid metabolism.

67.2  Identification of neuropathy target esterase 67.2.1  Organophosphorus Ester Insecticides: General Reactions with Serine Esterases Serine esterases belong to a larger family of proteins known as / hydrolases (Bourne et al., 1999; Taylor, 1992), whose members possess a similar fold with the nucleophilic residue residing on a flexible loop known as a “nucleophilic elbow” (Nardini and Dijkstra, 1999). AChE, the intended target for OP insecticides (Thompson and Richardson, 2004), is a serine esterase whose nucleophilic serine residue is rendered reactive by the nearby presence of a histidine and a glutamate, which together form the components of a classic Ser-His-Asp/Glu catalytic triad in the active site of the enzyme. Serine esterases catalyze the hydrolysis of carboxylate esters by the mechanism shown, in highly simplified form, in Figure 67.1A, which involves formation of a covalent acyl enzyme intermediate. This mechanism is essentially the same in serine proteases such as chymotrypsin, which cleave peptide bonds with intermediate formation of a covalent acyl enzyme (Aldridge and Reiner, 1972). OP insecticides have been made from a variety of derivatized or underivatized subclasses of OP compounds, including phosphates, phosphonates, phosphinates, and phosphoramidates (Thompson and Richardson, 2004).

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Figure 67.1  Reaction of a serine esterase with a carboxylate ester substrate and with organophosphorus inhibitors. The reactive serine residue at the enzyme active site is represented by -OH. (A) Following activation via abstraction of its hydrogen, the serine residue makes a nucleophilic attack on the acyl carbon of the ester and forms a tetrahedral intermediate. The alcohol moiety is rapidly expelled from this intermediate to produce a covalent acyl enzyme. Rapid aqueous hydrolysis of the acyl enzyme liberates the carboxylic acid and regenerates free enzyme. (B) Part of the efficacy of OP esters as serine esterase inhibitors results from their structural resemblance to the tetrahedral intermediate formed between the enzyme (NTE) and the carboxylate ester substrate. The rate of hydrolysis/reactivation of the organophosphorylated or organophosphinylated enzyme is much slower than that of the acyl enzyme, resulting in essentially irreversible inhibition. In addition, neuropathic organophosphorus compounds, such as certain organophosphonates (shown in the figure) and organophosphates, are able to undergo a second reaction termed “aging.” This entails loss of one of the R groups from the organophosphorylated enzyme leaving a negatively charged species attached to the active site and is thought to be the initiating step for OPIDN. (C) On the other hand, although the initial stage of NTE inhibition by organophosphinates resembles that of organophosphates and organophosphonates, these compounds are unable to undergo the aging reaction and are therefore unable to initiate OPIDN (Aldridge and Reiner, 1972; Richardson et al., 2009).

Overall, these compounds are hydrolyzable esters and act as pseudo-substrates for a variety of serine esterases and proteases. As an example, the reaction of an organophosphonate with a serine esterase is shown in simplified form in Figure 67.1B. The rate of hydrolysis of the phosphonylated enzyme is greatly (6–10 orders of magnitude) reduced

compared to that of the acyl enzyme. Thus, the enzyme becomes virtually permanently inhibited, although certain nucleophilic agents, such as oximes or fluoride anion, can catalyze a speedier dephosphonylation – hence the therapeutic use of oximes in the treatment of poisoning by anti-AChE OP compounds. However, the phosphonylated

Chapter | 67  Neuropathy Target Esterase

enzyme can subsequently undergo a second reaction, known as “aging,” which, in the case shown of an organophosphonate, results in the net liberation of the R group that had been attached to phosphorus via an oxygen atom (see Figure 67.1B). This dealkylation leaves the active site serine covalently attached to a negatively charged organophosphonyl moiety, which is significantly more resistant than the nonaged organophosphonyl group to removal by therapeutic nucleophiles. Organophosphinates also covalently react with the active site serine but cannot undergo the aging reaction because both R groups are attached to phosphorus by stable C–P bonds (see Figure 67.1C). More detail on these reactions and their associated kinetics can be found in other sources (Aldridge and Reiner, 1972; Richardson, 1992; Richardson et al., 2009)

67.2.2  Neuropathy Target Esterase as the Target for Initiation of Organophosphorus Ester-Induced Delayed Neuropathy Human NTE (UniProtKB/Swiss-Prot Q8IY17; PLPL6_ HUMAN) is a membrane-bound protein that is now considered to be lysophospholipase (EC 3.1.1.5) because it catalyzes the conversion of phosphatidylcholine to glycero­ phosphocholine (see the UniProt website http://www. uniprot.org/uniprot/Q8IY17). It is also cataloged as patatinlike phospholipase domain-containing protein-6 (gene name PNPLA6) because the catalytic domain of NTE contains a region homologous with patatin, a lipid acyl hydrolase found in potatoes and other plants. Although NTE can function biochemically as a phospholipase/lysophospho­ lipase, its precise physiological and pathogenic roles have yet to be established conclusively (see Section 67.4). Long before its present molecular classification, NTE was identified in a search for the target of neuropathic OP compounds. Elucidation of the events that initiate OPIDN involved a sequence of observations in vitro and ex vivo on the interaction of radiolabeled diisopropyl phosphorofluoridate (DFP) and other (unlabeled) serine esterase inhibitors (OP compounds, carbamates, and sulfonyl fluorides) with homogenized whole brain of adult chickens (usually hens). The data and interpretations have been detailed by Johnson (1990). In brief, because preliminary study showed that DFP, an OP inhibitor of many serine esterases that produces both cholinergic toxicity and OPIDN, became covalently bound only to proteins, criteria were adopted for a putative target site that would be a protein able to undergo the same set of general reactions shown in Figure 67.1. It was found that approximately 5% of the total DFP-labeling sites in hen brain were not covalently blocked by a variety of serine esterase inhibitors known not to cause OPIDN but were inhibited at toxicologically relevant doses by neuropathic OP compounds. Extensive screening then showed that only 2 out of more than 60 hydrolyzable esters, lipids, and/or

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peptides competed with labeling of that subset, suggesting that only these 2 had significant affinity for the relevant DFP-reactive sites. Finally, only one of several esterases that could hydrolyze these competitors (phenyl esters of phenylacetic acid and valeric acid) shared the same inhibitor responses as the apparently homogeneous target site (Johnson, 1969a,b, 1970). This esterase was dubbed NTE for neurotoxic esterase and, later, neuropathy target esterase; its ability to hydrolyze phenyl valerate (PV) in a reaction sensitive to neuropathic OP inhibitors forms the basis for a widely used in vitro screening test for such OP compounds (Kayyali et al., 1991). A threshold level of approximately 70–90% inhibition of NTE activity in the brain and spinal cord of test adult chickens has been found to be the norm for precipitation of clinically visible OPIDN, and this criterion has been incorporated into current regulatory guidelines [Organization for Economic Cooperation and Development (OECD), 1995; U.S Environmental Protection Agency (EPA), 1991; see Section 67.3]. Progressive inhibitors of NTE (i.e., those producing increasing inactivation with time; Aldridge and Reiner, 1972) were found to fall into two toxicological classes: neuropathic and nonneuropathic (Richardson, 2005). Neuropathic inhibitors were shown to be OP compounds of the phosphate, phosphonate, or phosphoramidate chemical classes, which are all capable of undergoing an aging reaction analogous to that shown in Figure 67.1B. Nonneuropathic inhibitors belonged to chemical classes that were incapable of undergoing the aging reaction. However, nonneuropathic compounds were found to be far from biologically inert: they were actually specifically prophylactic against OPIDN (not against acute anticholinesterase effects) if given to chickens prior to a neuropathic OP compound. These prophylactic compounds included certain carbamates (Johnson and Lauwerys, 1969), sulfonyl fluorides (Johnson, 1970), and phosphinates (as shown in Figure 67.1C; Johnson, 1974). The most striking evidence for the validity of NTE as the true target for initiation of OPIDN has been the correlation in time of the degree of inhibition of NTE (and of the radiolabeled target) by the nonaging compounds with their prophylactic effect. Both short-term (hours) and long-term (up to 5 days) prophylaxis are possible according to the structure of the agent, and the “window” of protection correlates with the persistence of inhibition until approximately 70–90% of NTE is again available to neuropathic challenge by virtue of reactivation or turnover of the inhibited enzyme (summarized by Johnson, 1990). Induction of OPIDN by aging NTE inhibitors and its prophylaxis by nonaging NTE inhibitors led to the proposal that generation of a negative charge at the active site of NTE was critical for initiation of the disease (Atkins and Glynn, 2000; Johnson, 1974). An elegant confirmation of the importance of the aging reaction in initiation of OPIDN was the demonstration that in the case of chiral phosphonates such as EPN oxon (O-ethyl O-4-nitrophenyl

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phenylphosphonate), both enantiomers can inhibit NTE at tolerated doses in vivo. However, the one that engages in the subsequent aging reaction produces OPIDN, whereas the other nonaging isomer is prophylactic (Johnson and Read, 1987). This result is obviously important mechanistically, but it also raises special questions about the interpretation of regulatory tests for chiral compounds (see Section 67.3.3.3).

67.2.3  Possible Involvement of Other Esterases in Organophosphorus EsterInduced Delayed Neuropathy Research results accumulated during past decades have supported the early identification of membrane-bound NTE as the site for initiation of OPIDN. However, although all detectable binding sites for [32P]-labeled DFP in hen brain were apparently dissected in Johnson’s seminal work (Johnson, 1969a,b), it has been proposed that further sites were actually present at such low concentration that they were undetectable but might have equal claim to be the initiating site. No convincing evidence for this idea has emerged over many years. However, the presence of a soluble apparent isoform of NTE activity has been reported by Vilanova and colleagues (Garcia-Pérez et al., 2003; Vilanova et al., 1999). NTE is only one of at least six distinct phenyl valerate hydrolases easily demonstrated to be present in hen brain (Johnson, 1982a), and until recently, no association of any of these other esterases with normal or abnormal processes had been demonstrated. However, Vilanova and colleagues have made extensive studies of NTE and related enzymes in sciatic nerve of the hen because, although this tissue has far less NTE activity than brain, it contains long axons that undergo degenerative changes in OPIDN (Vilanova et al., 1999). They have identified and studied what appears to be a freely soluble NTE-like enzyme according to inhibition characteristics. Apart from having a lower apparent molecular size than NTE from brain particles (Escudero and Vilanova, 1997), the most striking characteristic of this and related soluble sciatic nerve PV hydrolases is the sensitivity of one or more of these esterases to paraoxon. They react rapidly and progressively with this inhibitor to form covalently inhibited enzymes that spontaneously reactivate within a few hours at body temperature. Thus, in the standard in vitro laboratory assay for NTE, such enzymes are excluded by the paraoxon preincubation, but in vivo they would, for practical purposes, appear to be insensitive to that OP inhibitor and could therefore be added to the list of putative targets for the site of initiation of OPIDN (Barril et al., 1999). The uncovering of these enzymes is a tribute to a more careful kinetic analysis of the OP–enzyme interaction than has been possible in some general screening operations. Furthermore,

Hayes’ Handbook of Pesticide Toxicology

these researchers have shown that paraoxon competes strongly with substrate in the assay of soluble NTE so that the low activity in sciatic nerve has routinely been underestimated by 20-fold or more (Barril and Vilanova, 1997). Fortunately, neither of these confounding factors seems to exist with regard to particulate NTE. The possible relevance of any of these soluble enzymes to either initiation of OPIDN or its potentiation (elicitation of OPIDN by nonneuropathic NTE inhibitors after prior treatment with a near-threshold dose of a neuropathic NTE inhibitor – a phenomenon discussed in Chapter 69, in which it is called promotion) (Randall et al., 1997) requires evaluation by structure–activity relationship (SAR) studies using tissue from hens dosed with a battery of neuropathic and nonneuropathic OP compounds. To date, studies in vitro with the soluble enzymes (Vilanova et al., 1999) discourage an association because sensitivity to mipafox in vitro is so considerable (20-min I50  0.1 M compared with 7 M for NTE) that one might expect that such activities would be fully inhibited in vivo at doses that would scarcely affect NTE and would be well below the neuropathic or promoting dose. Lotti and colleagues have produced some correlative SAR evidence to suggest that the target for promotion of OPIDN by certain OP compounds is not one of the two paraoxon-sensitive hydrolases first identified by Poulsen and Aldridge (1964). Instead, the target may be a component of the particulate PV hydrolases, which they dubbed M-200 based on its comparatively low sensitivity to mipafox. M-200 is a portion of the C activity (see Section 67.4.2), which can be segregated in assays in which the mipafox concentration is raised (from the 50 M used in standard assays) to detect an enzyme with a 20-min I50 of 200 M (Lotti and Moretto, 1999; Milatovic et al., 1997). Vilanova and colleagues have also identified a component of soluble PV hydrolases that fits the characteristics of the target for promoters rather than initiators of OPIDN (Céspedes et al., 1997; Vilanova et al., 1999). However, although it seems likely that the target for promotion of OPIDN or other axonopathies is a serine esterase distinct from NTE, subsequent efforts to identify it have not been successful (Moretto et al., 2005, 2007). Research employing bioinformatic and functional proteomic approaches has revealed that more than 1000 serine hydrolases are encoded by the human genome (Casida et al., 2008). Of these, many are lipases and sensitive to inhibition by certain OP compounds. NTE has emerged as a probable lysophospholipase that hydrolyzes lysophosphatidylcholine, and the inhibitor profile of NTE indicates that it remains the most likely target for initiation of OPIDN. Nevertheless, elucidating the overall pathogenic mechanism(s) of this disease may depend on achieving a more complete understanding of other related lysophospholipases and the homeostasis of their substrates in target neural tissue (Vose et al., 2008), as discussed in Section 67.6.

Chapter | 67  Neuropathy Target Esterase

67.3  Toxicological applications 67.3.1  Hen Test As noted in Chapter 69, OP pesticides are screened for their relative abilities to inhibit AChE and to cause OPIDN; consequently, humans and susceptible animals are unlikely to develop OPIDN without acute cholinergic toxicity following OP pesticide exposure. The hen test has developed in sophistication from its origins in studies in the early 1930s, which showed that mature adult chickens were the most satisfactory animals to detect the OPIDN potential of various OP compounds found in the toxic “Ginger Jake” drink that paralyzed thousands of people in the southern United States (Smith et al., 1930). OPIDN in the adult chicken is the best available model for the human syndrome. Historically, it was found that hens responded positively and uniformly in tests of compounds suspected of causing OPIDN in humans (Davis and Richardson, 1980). The best-known examples of neuropathic compounds were tricresyl phosphates (not insecticides) as well as mipafox and leptophos (a pesticide candidate and pesticide, respectively) that were responsible for OPIDN incidents in the United Kingdom and the United States. The spinal tracts known to suffer selective damage in human OPIDN are likewise affected in hens, and damage to these regions is easily detected by standard histopathological techniques (see Chapter 69). Furthermore, slight abnormalities of gait are easily detected in these bipeds. Biochemically, the sensitivity of hen and human NTE to inhibition by OP inhibitors is similar (Lotti and Johnson, 1978). Since the 1950s, the UK Ministry of Agriculture Fisheries and Food has required information about the OPIDN potential of OP pesticides be submitted for registration. Briefly, batches of hens are required to be dosed with approximately the maximum tolerated dose accompanied by therapeutic measures to carry them through the inevitable cholinergic crisis. After observation over a period of 21 days for signs of ataxia, positively affected birds are necropsied and histopathological signs are sought. Unaffected birds are redosed and the observation is continued for an additional 21 days, with sample birds being necropsied at the conclusion regardless of the presence or absence of clinical signs. Since the 1970s, the U.S. EPA has required clinical and histopathological tests after long-term (often 90 days) feeding of tolerated low levels of OP compounds. However, despite the fact that such testing procedures are mandatory, reports continue to appear of occasional cases of OPIDN resulting from high-level occupational exposure or suicide attempts with pesticides, including methamidophos, leptophos, dichlorvos, trichlorphon, chlorpyrifos, and EPN (Lotti, 1992). The possibility of less severe clinical cases being overlooked has caused concern, and improvements in the discriminatory power of regulatory tests for OPIDN are clearly desirable.

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Assay of NTE activity in appropriate necropsy samples taken soon after dosing first became an optional extra, then recommended, and finally a required component of OPIDN toxicity tests over a period of approximately 20 years following the first reports of the enzyme in 1969, although some manufacturers adopted the procedure voluntarily very early (OECD, 1995; U.S. EPA, 1991). The biochemical test does not replace clinical and histopathological observations but, rather, complements them (European Centre for Ecotoxicology and Toxicology of Chemicals, 1998; Johnson, 1984). It has the advantage that unlike the subjective and qualitative tests, it is quantitative so that the degree of risk of OPIDN arising from a defined dose can be assessed; hitherto the conclusion from tests could only be indicated as “yes/no” or occasionally “marginal.” Furthermore, sufficient data on the relationship of chemical structure to neuropathic response have accumulated in experimental studies in vivo (Johnson, 1975a) and in vitro (the latter using both human and animal tissues; Johnson, 1975b, 1988; Lotti and Johnson, 1978) that fairly confident predictions of the OPIDN potential of an untested compound can be made (Davis et al., 1985). Monitoring of NTE activity in accessible tissue samples taken from a few patients who have deliberately ingested known OP compounds confirms the relationship (Moretto and Lotti, 1998). Although the early work studied brain tissue only, it is accepted that clinical signs of OPIDN reflect the fact that neuropathic lesions are scarce in brain and more marked in spinal cord and peripheral nerve (see Chapter 69). In some early experiments with dichlorvos (Johnson, 1978), the dose (although many times the LD50) appeared not to have reached the spinal cord sufficiently to inhibit NTE and no clinical signs developed in pair-dosed birds, although brain NTE was inhibited. A further dose was necessary to increase inhibition in spinal cord and to precipitate clinical neuropathy. For this reason, it is now customary for ex vivo assays from dosed birds to study both tissues and it is not uncommon to find slightly less inhibition in spinal cord than in brain: the threshold figure of inhibition is accordingly set slightly lower for spinal cord.

67.3.2  Structure–Activity Relationships   and Prediction of Organophosphorus   Ester-Induced Delayed Neuropathy   Potential in Hens The ability to analyze both positive and negative clinical responses in terms of the degree of effect on NTE during OPIDN tests made it possible to review a large amount of test data and to define guidelines for predicting neuropathic potential of both the plasticizer and the pesticide types of OP esters (Davis et al., 1985; Johnson, 1975a, 1982a). Guidelines (to be taken in concert) concerning the likelihood that a pesticide-type OP ester with general structure

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R1R2P(O or S)X will cause neuropathy at less than lethal doses are as follows: 1. Factors that increase OPIDN potential more than acute toxicity include: a. Choice of phosphonates or phosphoramidates rather than analogous phosphates b. Increase in chain length or hydrophobicity of R1 and R2 c. A leaving group X that does not sterically hinder approach to the active site of NTE. 2. Factors that decrease the comparative potential include: a. The converse of factors 1a-c b. Choice of R or X groups that are very bulky (e.g., naphthyloxy) or nonplanar c. Choice of a nitrophenyl group at X d. Choice of comparatively more hydrophilic X groups (oximes or heterocyclics) e. Choice of thioether linkages at X. Although inclusion of NTE assays in toxicological tests is now mandatory for all OP pesticides submitted for registration or re-registration, most data are not published in the open literature. Considering the structure–activity factors, it is clear why malathion, parathion, fenitrothion, and diazinon (compounds with a higher affinity for inhibiting AChE as opposed to NTE) are among the compounds listed as far below the hazard line for OPIDN and why EPN, a phosphono­thioate with a hydrophobic phenyl group at R, is neuropathic even with a 4-nitrophenyl leaving group. In addition, it is not surprising that other phenylphosphono­ thioates, such as desbromo-leptophos or cyanofenphos, are also neuropathic (Johnson, 1975a; Soliman et al., 1986) and that, in its homologous series, only dichlorvos is not neuropathic at the LD50 dose. A part from the obvious correlation with clinical effects at both ends of the range of inhibition, there was a valuable warning of risk with chlorpyrifos dosed to hens at approximately the LD50: at that time, this pesticide was regarded as nonneuropathic based on approved tests involving only clinical and histopathological measures. The risk indication was vindicated by positive OPIDN seen both in hens given higher doses and in a failed human suicide attempt in which death was averted through medical intervention (Capodicasa et al., 1991; Moretto and Lotti, 1998). However, biochemical studies of the type discussed in the following section supported the fact that chlorpyrifos could not produce OPIDN at doses below the LD50 (Richardson et al., 1993a).

67.3.3  Application of Neuropathy Target Esterase Studies to Human Risk Assessment 67.3.3.1  In Vitro Comparison of Enzyme Targets When OP esters that exert biological effects in vivo are administered to animals, they are subjected to a variety of

processes (e.g., absorption, distribution, metabolic activation and/or deactivation, and excretion) before a certain amount of a proximal toxicant is delivered to the ultimate targets: AChE in the case of acute toxicity and NTE for delayed neuropathy. However, because the targets for both acute and delayed neurotoxicity are associated with nervous tissue, it seems reasonable to propose that whatever percentage of a dose ultimately reached the nervous system in appropriately reactive form, that amount would prefer to react with AChE or NTE according to the relative potencies demonstrable with these enzymes in vitro. Use of the in vitro ratios, which can be determined easily and very early in a development program, may serve to guide synthetic chemists away from structures carrying neuropathic hazard. In recent work (Kropp and Richardson, 2003; Richardson et al., 2009), the in vitro ratio has been redefined as the reciprocal of the original ratio proposed by Lotti and Johnson (1978). Thus, we now use the relative inhibitory potency (RIP), which is defined by Eq. (1), where ki is the bimolecular rate constant of inhibition under specified conditions (e.g., temperature, pH, and ionic strength):

RIP 5 ki (AChE)/ki (NTE)

(1)

Although it is better to use kinetic methods for determining inhibitory potency, fixed-time inhibition has also been used directly to obtain the I50, which is the inhibitor concentration required to produce 50% inhibition of the enzyme at specified conditions (e.g., time, temperature, pH, and ionic strength). Because inhibition of serine esterases by OP inhibitors is time dependent, the time of pre­ incubation of inhibitor must be given with the I50. Many of the I50 values reported in the literature have been standardized on a 20-min preincubation interval. Given that the I50 is reciprocally related to the ki as shown in Eq. (2), where t is the time of preincubation of inhibitor and enzyme, then an alternative form of the RIP in terms of I50 values is given by Eq. (3):

I 50 5 ln 2 /(ki  t )

(2)



RIP 5 I 50 (NTE)/I 50 (AChE)

(3)

When the RIP is greater than 1, the potency of the OP inhibitor is greater for AChE than for NTE, and cholinergic toxicity is favored over delayed neurotoxicity. We have found that under these conditions, doses greater than the LD50 are required to produce neuropathy with most compounds; obviously, such large doses of an anticholinesterase agent would require prophylaxis against the cholinergic effects in order to enable the animals to survive so that delayed effects could be seen.

Chapter | 67  Neuropathy Target Esterase

Although such in vitro/in vivo correlations have been obtained, the following assumptions built into the system need to be recognized: 1. For compounds that require metabolic activation in vivo, the actual inhibitor species must be identified and tested in vitro. 2. The inhibitor has equal access to both enzymes in the brain. 3. The rates of synthesis of fresh NTE and AChE are sufficiently slow so as not to affect the prognosis after an intoxication that causes massive inhibition of enzymes. 4. The rate of aging of NTE is rapid compared with the rates of synthesis of new enzymes. 5. The extent of spontaneous reactivation of inhibited enzymes in vivo is small. 6. The hen is an adequate model for humans. Assumption 1 is often possible for OP pesticides. There is good evidence to support assumptions 2–4. Assumption 5 appears to be true for NTE in our experience; inhibited NTE ages with a half-life of less than 1 h to a nonreactivatable form in most cases tested, and there is little spontaneous reactivation of NTE. However, assumption 5 is not true in all cases of AChE inhibition. For example, after poisoning with haloxon, the di-2-chloroethylphosphorylated AChE has a half-life of approximately 22 min. Consequently, the acute toxicity of haloxon is less than might have been predicted. Thus, although the ratio of I50 values for this compound is approximately 100, it is in fact neuropathic at less than the LD50 dose (Johnson, 1982a). For a similar reason, methamidophos has caused OPIDN in humans, although in hens neuropathy is not caused by doses less than six to eight times the unprotected LD50 given with aggressive treatment to prevent cholinergic death (Johnson, 1981; Lotti, 1992; Senanayake and Johnson, 1982). Assumption 6 is considered in more detail later. The previous considerations led to the suggestion that any compound for which the RIP is less than 20 should be viewed with strong suspicion and that neuropathy may be caused in atropinized birds with single doses of some compounds where the ratio is as high as 100. Despite such a large difference in species type, it has been shown that the target enzymes in humans and hens are similar in their response to OP inhibitors; the I50 values for both AChE and NTE differed between species by no more than a factor of 4 and often by less (Lotti and Johnson, 1978). However, these variations were not identical for each enzyme so that ratios of I50 values diverged up to eightfold in some cases. For cases in which the ratio is higher for humans, one might predict it would be comparatively more difficult to produce neuropathy in humans than in hens. However, in the case of trichlorphon (dichlorvos being the active inhibitory species), the in vitro ratio is three times lower for humans than for hens. There is increasing evidence that trichlorphon in single massive

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doses can produce neuropathy in humans, whereas it requires more than one dose in hens (Hierons and Johnson, 1978; Johnson, 1981). This may reflect pharmacokinetic differences, the greater relative sensitivity of the human target, or the fact that a lower proportion of NTE in the human nervous system might be required to be phosphorylated and aged to reach the initiation threshold. Although hen brain has most often been used as the source of NTE and AChE for in vitro comparisons of these enzyme targets to inhibition by direct-acting OP inhibitors, mice and rats have these enzymes as well (Novak and Padilla, 1986; Veronesi et al., 1986, 1991). Therefore, these species should be considered as models for in vitro or ex vivo evaluations, despite the relative lack of a robust clinical response to neuropathic OP compounds in rodents.

67.3.3.2  Neuropathy Target Esterase and the Evaluation of Organophosphorus EsterInduced Delayed Neuropathy in Humans Lotti (1992) listed 10 different pesticides that have been reported to cause OPIDN in humans, and isofenphos has subsequently been added to the list (Moretto and Lotti, 1998). According to the SAR studies discussed previously, 9 of these have molecular structures indicative of significant neuropathic potential, which has also been confirmed in laboratory studies involving assay of NTE in hen necropsy samples: Lotti noted that a case report naming omethoate (not significantly anti-NTE) as a causative agent had no sound evidence to identify the poison. A case of OPIDN believed to be due to a suicide attempt with a massive dose of parathion (De Jager et al., 1981) was highly unusual in that it appeared to be a solitary event even though acute poisoning with this pesticide is probably the most common of all reported OP intoxications. Parathion and paraoxon have very low anti-NTE potential, but the oxon derived from the impurity ethyl bis(4-nitrophenyl) phosphorothioate is a potent inhibitor (Johnson, 1982b) and this could well have been the actual causative agent. In addition, in clinical reports depicting individual cases of delayed neurotoxicity arising from carbamate ingestion (Dickoff et al., 1987; Umehara et al., 1991), these may be due to similar contamination of the ingested pesticide. The threshold of NTE inhibition that precipitates OPIDN in humans might be established if the following steps were taken more often following severe poisonings due to OP compounds: 1. The actual agent involved should be identified by chemical analysis. 2. The sample should be analyzed for major and minor OP constituents. 3. Lymphocyte NTE as well as erythrocyte AChE should be assayed in (serial) blood samples during treatment (see Section 67.3.4).

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4. In the event of fatal poisoning, immediate autopsy samples should be obtained from brain and spinal cord and deep frozen until AChE and NTE assays can be performed. A few such investigations have been performed with patients who did and others who did not suffer a neuropathic consequence of severe OP poisoning (Lotti et al., 1981; Moretto and Lotti, 1998). The authors concluded that only substantial peak inhibition of NTE was associated with expression of frank clinical OPIDN. Based on the preceding limited information, the application of the in vitro ratios of enzyme sensitivity appears to be an acceptable procedure in predictive toxicology and in clinical prognosis.

67.3.3.3  Neuropathy Target Esterase and the Assessment of Chiral Compounds One unresolved toxicological issue is the problem of chiral compounds (Battershill et al., 2004; Johnson, 1987). The relative sensitivities of AChE and NTE in humans and hens to most tested OP compounds are not greatly different so that the measured relative susceptibility of hens to cholinergic or OPIDN effects can be transposed to humans. The actual dose effective in these species may differ because, in general, mammals have greater capacity for both bioactivation and detoxification of chemicals, but this does not alter the relative anti-esterase activities of these products. However, chiral OP compounds are usually a 50/50 molar mixture of two distinct chemicals that may be metabolized differently and to different extents in hens (Johnson et al., 1991) and differently yet again in humans. Furthermore, the relative anti-AChE/anti-NTE potencies of the various metabolic products are unlikely to be all the same. Thus, in a worst-case scenario, the predominant form of the anti-esterase compound(s) circulating in a dosed hen may dominantly affect AChE, whereas a different (anti-NTE) compound(s) might predominate in humans. Although the problems of doing full toxicological evaluations of resolved isomers would be immense, the following comparatively easy investigations could be useful: 1. Perform assays of AChE and NTE in necropsy samples from a few mammals (rats or mice) dosed with racemic compound and run in parallel with the full OPIDN evaluation in hens. Neither clinical nor histopathological examination is needed because the objective is to decide whether the mammal produces markedly different relative effects on the enzyme targets than does the hen. 2. Run assays in vitro for the potency against AChE and NTE of resolved isomers of whatever anti-esterase compounds have been identified during routine metabol­ ism studies of the unresolved compound: these assays are straightforward and require only a very few milligrams of material, whereas whole-animal studies might require many grams.

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Taken together, performing these suggested limited studies should indicate whether the hen study with racemic compound is indicative of the human situation. Moreover, a similar strategy could be used for nonchiral compounds in early stages of product development when there would likely be limited amounts of compounds for testing.

67.3.3.4  Neuropathy Target Esterase and   the Assessment of Effects of Long-Term,   Low-Level Exposure A significant concern is whether short-term, high-dose experiments in hens are appropriate to assess the possible neuropathic hazard of long-term human exposure to relatively lower levels of OP pesticides, with the possibility of cumulative effects. The quantitative data on inhibition of NTE that emerge after short-term tests, even when no other effects are seen, go some way in providing a useful assessment. Thus, whereas a single dose of 50 mg/kg of mono-ocresyl diphenyl phosphate caused OPIDN in hens, a total of 175 mg/kg dosed daily over 10 weeks did not. Monitoring brain and spinal cord NTE activity over this period showed that inhibition reached a steady state after 1 or 2 weeks at a value (45–60%) that is below the threshold (7090%) required to initiate OPIDN (Lotti and Johnson, 1980a). Similarly, whereas a single oral dose of 300 mg/kg of chlorpyrifos in atropinized hens produced 77% brain NTE inhibition (Richardson et al., 1993a), daily oral dosing of 10 mg/kg/day chlorpyrifos for 20 days produced brain NTE inhibition that fluctuated between 1 and 18% during the dosing interval, with no inhibition 4 weeks thereafter and no clinical signs of neuropathy throughout the experiment (Richardson et al., 1993b). OECD Guideline 419 restricts multidose tests to 28 days unless special exposure conditions pertain and suggests that negative results on the biochemical, histopathological, and behavioral endpoints indicate that further testing of the compound is not required (OECD, 1995).

67.3.4  Biomarkers and Biosensors The topic of biomarkers and biosensors of exposure to neuropathic OP compounds has been extensively reviewed (Richardson et al., 2009) and is discussed only briefly here. It is well-known that erythrocyte AChE and serum or plasma butyrylcholinesterase (BChE) can be used as biomarkers of exposure to anticholinesterases, such as OP insecticides (Lotti, 1995; Wilson and Henderson, 1992). It would be desirable to have a blood biomarker for potentially neuropathic OP compounds, but erythrocytes as well as serum or plasma do not contain NTE. Fortunately, lymphocytes do contain NTE activity, which was discovered by Dudek and Richardson (1978, 1982). Animal studies have been used to test correlations of inhibition in lymphocytes with that in brain (Makhaeva

Chapter | 67  Neuropathy Target Esterase

et al., 2003, 2007; Richardson and Dudek, 1983; Richardson et al., 1993a,b; Schwab and Richardson, 1986). Lymphocyte NTE activity has been characterized in human populations (Bertoncin et al., 1985; Maroni and Bleecker, 1986) and used to monitor occupational exposures to potentially neuropathic OP compounds (Lotti et al., 1983, 1986; McConnell et al., 1999). These studies show that NTE in lymphocytes or whole blood can be assayed and used as a biomarker of exposure to neuropathic OP compounds, particularly if the blood samples are taken within 24 h of acute exposures. Apparently, there have been no attempts to measure aging of NTE in lymphocytes of exposed animals or humans. However, given that aging of NTE inhibited by neuropathic OP compounds typically occurs with a half-life of a few minutes (Clothier and Johnson, 1979; Johnson, 1982a), aging would be complete by the time that blood samples could be taken, processed, and assayed for activity. Nevertheless, given the apparent requirement for aging as well as inhibition of NTE in OPIDN, to help rule out false positives arising from inhibition by nonaging inhibitors, aging studies should be undertaken in future work. Biosensors that could be used for testing the neuropathic potential of OP compounds or detecting their presence in the environment have been constructed using the catalytic domain of NTE (Kohli et al., 2007). These biosensors can also incorporate AChE and BChE so that they could be used to discriminate between cholinergic or delayed neuropathic compounds as well as to carry out RIP studies (see Section 67.3.3.1).

67.4  Nature and properties of neuropathy target esterase 67.4.1  Biochemical Studies Until quite recently, all studies on NTE relied on detection of its esterase activity or labeling by [3H]DFP. Using these methods, it was shown that in the adult chicken, the highest specific activities of NTE were found in brain, whereas spinal cord and sciatic nerve contained substantially less (Dudek and Richardson, 1982; Johnson, 1982a). In dissected areas of human brain, NTE activity varied by a factor of 2, with cerebral cortex the highest and cerebellum the least (Lotti and Johnson, 1980b). Relatively high levels of NTE have been found in several nonneural chicken tissues, including intestine, spleen, and thymus (Johnson, 1982a), as well as circulating lymphocytes (Dudek and Richardson, 1982). Extremely high levels have been found in bovine adrenal medulla (Sogorb et al., 1994). Cultured bovine chromaffin cells and human neuroblastoma cell lines have been shown to have substantial NTE activity and have been suggested as in vitro systems for

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assessment of neuropathic OP compounds (Sogorb et al., 1997; Veronesi et al., 1997). Biochemical fractionation of chicken brain homoge­nates showed that NTE is enriched in microsomal membrane fractions (Richardson et al., 1979). [3H]DFP labeling indicated that NTE comprised less than 0.1% of total microsomal protein (Williams and Johnson, 1981). NTE is an integral membrane protein, as indicated by its requirement for detergent for solubilization and predictions based on its primary sequence (Wijeyesakere et al., 2007a). The type and concentration of detergent were shown to be important for the maintenance of NTE activity (Davis and Richardson, 1987); fractionation of the solubilized material generally led to substantial loss in NTE activity, which could be partially ameliorated by the addition of phospholipids (Pope and Padilla, 1989a). On sodium dodecyl sulfate polyacrylamide gel electrophoresis (SDS-PAGE), [3H]DFP-labeled NTE appears as a 155-kDa polypeptide (Williams and Johnson, 1981), whereas on gel filtration, detergent-solubilized NTE migrates as a complex with an apparent molecular weight greater than 850 kDa (Pope and Padilla, 1989b; Thomas et al., 1990).

67.4.2  Enzymology Enzymatically, NTE behaves as a typical B-esterase; that is, rather than hydrolyzing OP compounds as A-esterases (e.g., paraoxonases) do, it is inhibited by them (Aldridge and Reiner, 1972). Although findings suggest that NTE may be a lysophospholipase and demonstrate that it hydrolyzes lysophosphatidylcholine (hence its classification as EC 3.1.1.5) (Quistad and Casida, 2004; Quistad et al., 2003; Van Tienhoven et al., 2002; Vose et al., 2008; Zaccheo et al., 2004), the physiological function of the enzyme has not been identified conclusively. However, it is known that NTE rapidly hydrolyzes certain hydrophobic artificial substrates, of which PV has the best combination of sensitivity and specificity under appropriate assay conditions (Johnson, 1988). Even with PV, a differential assay is necessary to distinguish NTE from other esterases that can hydrolyze the same substrate but do not have the appropriate inhibition characteristics to qualify as a target for OPIDN. Thus, when total hydrolytic activity in the absence of inhibitors is termed A, the paraoxon-resistant activity is termed B, and the residual activity resistant to both paraoxon and mipafox (either together or in sequence) is C, then the activity of NTE is determined as B  C and specificity as the ratio (B  C)/B. In an assay of hen brain PV hydrolases, after preincubation of the tissue with paraoxon, the substrate specificity is approximately 65%, which allows quite accurate determination of NTE by the B  C calculation even when overall activity is low, as in some necropsy samples from birds dosed with neuropathic compounds.

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Extensive structure–activity studies for both substrates and inhibitors in vitro have been reported (Borhan et al., 1995; Johnson, 1975b, 1988; Quistad and Casida, 2004; Thomas et al., 1990). Although most of these identified substrates are more sensitive than PV (catalytic center activity up to two- or threefold greater), all were less than 50% specific. Therefore, for routine investigations, PV is accepted as widely tested and approved. Nevertheless, some of these alternatives may be useful for specific studies, such as for a partly purified enzyme free of interfering esterases, when tissue activity is very low, or for kinetic investigations that require the substrate to have a sufficiently low Km to ensure complete quench of any progressive inhibition at the instant of the addition of substrate. It is a fact that the apparent Km of PV for NTE is high (10 mM compared with 0.1 mM for phenyl phenylacetate, a substrate used in early studies but which lacked both sensitivity and specificity; Johnson, 1982a). The aging reaction (see Figure 67.1B) of DFP-inhibited NTE is very rapid, with a half-life of a few minutes, and the aged isopropyl group is quantitatively transferred to a putative covalent acceptor site, termed site Z, within NTE. This contrasts with a much slower rate of aging for DFP-inhibited cholinesterases, in which the aged isopropyl group is liberated into free solution (Clothier and Johnson, 1979). Yoshida et al. (1995) investigated the reaction of chicken brain NTE with tritiated octyl cyclic saligenin phosphonate; these authors report that only approximately 15% of the aged saligenin group is transferred to the putative site Z. The identity of the putative site Z in NTE is unknown, but clearly it is a residue that is in close proximity to the active site in the native folded protein. There is evidence from proteolysis of SDS-solubilized preparations of [3H]DFP-labeled chicken brain microsomes that site Z lies within 150 residues of the active site serine residue (Glynn et al., 1993). Although the rapidity of aging of covalently bound OP and the putative intramolecular transfer of alkyl groups appears to be a unique feature of NTE, it has been concluded that the generation of a negatively charged species at the active site, rather than the modification of site Z, is the critical event in initiation of OPIDN (Johnson, 1990). Further support for this conclusion has been provided by studies on the neuropathic phosphoramidate, mipafox. It was initially proposed on theoretical grounds that aging of mipafoxinhibited NTE involves deprotonation, rather than dealkylation, to leave an electronegative species attached to the active site serine (Richardson, 1995). This contention was subsequently demonstrated experimentally for the mipafoxinhibited NTE catalytic domain (Kropp et al., 2004).

67.4.3  Isolation and Localization The low abundance and apparent requirement for membrane lipid to maintain NTE activity impeded its isolation

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for several years. A fraction substantially enriched in [3H]DFP-labeled NTE, but far from homogeneous, was isolated from chicken brain (Rueffer-Turner et al., 1992). An apparently homogeneous NTE preparation from phospholipase A2-solubilized embryonic chicken brain had a specific activity approximately half that of the initial crude solubilized extract (Mackay et al., 1996). A breakthrough was finally achieved by the synthesis of a novel reagent, S9B [1-(saligenin cyclic phosphoro)-9-biotinyldiaminononane], for affinity purification of NTE (Glynn et al., 1994). S9B reacted rapidly and specifically with NTE in brain microsomes and resulted in the covalent attachment, via a long alkyl spacer, of a biotin molecule to the active site serine residue. Microsomal proteins, quantitatively solubilized by boiling in dilute SDS, were then subjected to affinity chromatography with avidin-Sepharose, which binds biotinylated polypeptides. S9B-labeled NTE was eluted from the avidin by boiling in SDS. Two polypeptides (possibly carboxylases) with endogenous covalent biotin prosthetic groups that coeluted from avidin-Sepharose with NTE were removed by subsequent preparative electrophoresis (Glynn et al., 1994). Isolated chicken NTE was digested with endoproteinase Glu-C and the resulting peptide fragments were resolved by SDS-PAGE. The N-terminal amino acid sequence of one of these fragments provided sufficient information to synthesize an 11-residue peptide, which was used to raise a rabbit antiserum to NTE (Glynn et al., 1998). An immunohistochemical survey of the chicken nervous system using this antiserum showed that NTE was present in essentially all neurons but was absent from glia. NTE immunostaining could not be detected in normal sciatic nerve but accumulated at the constriction site 8 h after nerve ligation, indicating that NTE undergoes fast axonal transport. NTE immunostaining filled neuronal cell bodies (except the nucleus) and sometimes extended into the proximal axon; this pattern, taken together with the biochemical data on NTE in microsomal fractions (Richardson et al., 1979), indicated that NTE is probably associated with the endoplasmic reticulum (ER). These immunostaining characteristics were not detectably altered in chickens 1 or 3 days after treatment with a neuropathic OP compound, suggesting that OP-modified NTE was neither grossly redistributed nor degraded faster than native NTE (Glynn et al., 1998). The distribution of NTE expression was visualized by -galactosidase reporter staining in Nte/ mice (Nte is the gene encoding the NTE protein) (Winrow et al., 2003). Developing mice showed robust expression in the lens and spinal cord. In adults, expression was noted in Leydig cells of the testis. In the brain, staining was found in the cerebral cortex, CA1–CA3 regions of the hippocampus, and the Purkinje cell layer of the cerebellum; staining in all brain regions appeared to be associated with neurons. Staining was absent in the dentate gyrus and striatum.

Chapter | 67  Neuropathy Target Esterase

67.4.4  Molecular Biology The N-terminal sequence of an endoproteinase Glu-C fragment of S9B-labeled pig brain NTE was found to be very similar to a human-expressed sequence tag cDNA; the latter was used to initiate screening of human brain cDNA libraries from which a full-length NTE cDNA clone was finally isolated (Lush et al., 1998). The NTE cDNA clone D16 encoded a polypeptide of 1327 amino acids, and analysis of this sequence with transmembrane prediction programs (PredictProtein and TMHMM) indicated the presence of a single N-terminal transmembrane segment (Figure 67.2) (Wijeyesakere et al., 2007a). Biochemical experiments indicated that the active site serine residue labeled by S9B lay between residues 955 and 1033, and interest was focused on Ser966, which resided within the lipase motif Gly-XxxSer-Xxx-Gly, common to all serine hydrolases (Lush et al., 1998). Ser966 was subsequently confirmed as the active site residue by [3H]DFP labeling and protease digestion of a recombinant form of NTE (Atkins and Glynn, 2000). Furthermore, human NTE is highly homologous (41% identical) to a Drosophila neuronal protein called “Swiss cheese” (SWS; Lush et al., 1998). The sws mutation results in glial hyperwrapping of neurons, which in turn leads to apoptotic death of both cell types; the name Swiss cheese derives from the vacuolated appearance of the mutant brains (Kretzschmar et al., 1997). It has been suggested that the SWS protein is involved in a cell signaling pathway, and attention has been drawn to the similarity between an N-terminal domain of SWS (also present in NTE) and the cyclic AMPbinding regulatory subunit of protein kinase A (Kretzschmar et al., 1997). In addition to sharing sequence homology, biochemical assays have shown that NTE-like phenyl valerate hydrolase activity is present in wild-type Drosophila but absent from sws mutants (Moser et al., 2000). Given that

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SWS is the Drosophila homologue of NTE, it is reasonable to hypothesize that, by analogy, NTE may mediate cell signaling in the developing vertebrate brain. In situ hybridization experiments on mouse embryos show that NTE mRNA is expressed in neurons from their earliest appearance in the developing nervous system (Moser et al., 2000). Sequence database searches revealed that NTE consists of an N-terminal transmembrane domain (TMD), three cyclic nucleotide binding (CNP) domains (termed CNP1–CNP3), and a catalytic domain that has 30% homology to the plant protein patatin [termed the NTE patatin homology domain (PNTE)] (Figure 67.3A). Subsequent molecular modeling undertaken by Wijeyesakere and colleagues revealed that the catalytic center of NTE consists of a Ser966-Asp1086 dyad similar to that seen in cytosolic phospholipase A2 and patatin (Wijeyesakere et al., 2007a). The unusual organization of NTE domains (TMD-CNP1-CNP2-CNP3-PNTE) suggests that the protein may be a member of a novel family that appears to comprise potential serine hydrolases. Furthermore, NTE shares homology with a number of polypeptides predicted from the sequencing of genomes of bacteria, yeast, and nematodes (Lush et al., 1998). In particular, a 200-amino acid domain, corresponding to NTE residues 910–1109 (encompassing PNTE), is highly conserved (29% identity between human NTE and the Escherichia coli homologue YCHK) and, notably, all the homologous proteins contain a serine residue in the same position as Ser966 of NTE (see Figure 67.3B). In view of its patatin domain and ability to hydrolyze lysophospholipids, NTE is currently classified as patatinlike phospholipase domain-containing protein-6, whose corresponding gene name is PNPLA6 (see the UniProt website http://www.uniprot.org/uniprot/Q8IY17). The human gene has three splice variants in the UniProt entry, and isoform1 is regarded as the canonical sequence. Nevertheless, for consistency of sequence numbering in this chapter, we use isoform-2, consisting of 1327 amino acid residues yielding a molecular weight of 146 kDa with no post-translational modifications.

67.4.5  Structure of the Catalytic Domain Figure 67.2  Predicted domain organization of NTE. NTE is a protein comprising 1327 amino acids with five (predicted) domains: an N-terminal transmembrane domain (TMD); three putative regulatory domains (CNP1–CNP3), which have some similarity to the cyclic nucleotide binding protein known as catabolite activator protein; and a C-terminal effector domain (the NTE patatin homology domain; PNTE), which contains the esterase activity of NTE. The active site serine (Ser966) lies within PNTE (Wijeyesakere et al., 2007a). NTE residues 727–1216 (encompassing the PNTE domain) have been cloned into a pET vector and expressed in E. coli with a short N-terminal (T7) tag sequence and a C-terminal His6 tag; this construct, dubbed NEST, has the OP-sensitive phenyl valerate hydrolase activity of full-length NTE (Atkins and Glynn, 2000).

Although it is known that NTE is able to hydrolyze carboxylic acid esters and lysophospholipids, its tertiary structure is yet to be determined. Attempts to solve its crystal structure have failed due to its large size and membrane association. Nonetheless, homology modeling is a powerful technique for gaining insight into the folding of its various domains. Analysis of the PNTE domain revealed it to be homologous to the plant protein patatin, whose crystal structure has been solved (Rydel et al., 2003). This allowed us to develop a homology model for PNTE (Wijeyesakere et al., 2007a), thereby suggesting that PNTE possesses a modified / hydrolase fold with a Ser966-Asp1086 dyad (Figure 67.4),

Figure 67.3  (A) Alignment of domains along with the amino acid sequences in the highly conserved patatin-homology domain of NTE (PNTE) with homologous proteins from various species. The Uniprot accession number for the amino acid sequence of each protein is shown within parentheses. (B) Amino acids (aa) 933–1099 of NTE (complete sequence 5 1327 aa) that comprise PNTE are aligned with homologous regions from SWS (Drosophila; 1425 aa), YOL4 (Caenorhabditis elegans; 1351 aa), YMF9 (Saccharomyces cerevisiae; 1679 aa), and YCHK (E. coli; 314 aa). Residues identical in at least four of the proteins are shown white on black with homologous residues shown white on grey. The positions of the active site serine and aspartate residues (corresponding to Ser966 and Asp1086, respectively, in human NTE) are shown black on grey and are indicated by arrows.

Chapter | 67  Neuropathy Target Esterase

Figure 67.4  Homology model of the NTE patatin homology domain (PNTE). The Ser966-Asp1086 residues of the catalytic dyad are rendered as sticks (indicated by an arrow) (Wijeyesakere et al., 2007a).

as opposed to the previously hypothesized Ser966-Asp960Asp1086 catalytic triad (Atkins and Glynn, 2000). Thus, the PNTE model is similar to the overall folding and catalytic mechanism seen in other proteins, such as mammalian cytosolic phospholipase A2 (Dessen et al., 1999).

67.5  Role of neuropathy target esterase in organophosphorus ester-induced delayed neuropathy: a toxic gain of function? Although inhibition of NTE activity suggests a loss-offunction mechanism, the requirement for aging suggests otherwise. Whereas aging might be thought to render the enzyme irreversibly inhibited with no possibility of reactivation, sustained inhibition above the 70% threshold by repeated dosing of hens with nonaging inhibitors does not produce OPIDN (Johnson, 1970, 1974, 1982a). In addition, repeated administration of aging inhibitors to hens at doses that produce a continuous subthreshold level of inhibition does not produce OPIDN unless a higher single dose is given that pushes the inhibition over the threshold. In this

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case, a delay of at least 10 days from the spike above threshold is still required before clinical effects are seen (Johnson and Lotti, 1980; Lotti and Johnson, 1980a). These results, in concert with the structure–activity studies on aging and nonaging inhibitors, indicate that although inhibition is necessary for OPIDN, it is not sufficient; that is, a suprathreshold level of aging must also occur. Thus, whereas inhibition alone would represent a loss of function of NTE, aging would confer a gain of function (Glynn, 2000). Just how a gain of function could occur is not clear, but one theory is that the negatively charged phosphoryl group extrinsically added to NTE through inhibition and aging resembles the negatively charged phosphate added to proteins by protein kinases (Richardson, 1984). Such an adventitious phosphoprotein could trigger signaling events that would be deleterious to the neuron and produce axonal degeneration, perhaps through a putative axonal self-destruct program that has been proposed to be an important mechanism in neurodegenerative disorders such as motor neuron disease (Raff et al., 2002). To date, the phosphoprotein hypothesis of NTE aging has not been adequately tested experimentally. A fundamental tenet of biology is that structure dictates function. Thus, we would expect that if the aging of NTE produced a toxic gain of function, there would be an associated change in structure of the molecule. To begin to examine this possibility, we used our three-dimensional model of the patatin domain of NTE (PNTE) to undertake in silico inhibition and aging using the neuropathic OP compound DFP (Wijeyesakere et al., 2007b). The results showed that the backbone structures of the native, inhibited, and aged models were not significantly different from each other. We then experimentally determined the x-ray crystal structure of patatin that had been inhibited and aged with DFP (Wijeyesakere et al., 2009), and we found no significant change in backbone structure from that of the native protein (Rydel et al., 2003). Although this work needs to be extended to experimental studies of full-length NTE, if aging does not produce measurable structural changes in the molecule, it would be difficult to envision how the toxic gainof-function theory could be correct. However, it must be borne in mind that subtle changes in local structure within a macromolecule can have profound effects on function. For example, the much stronger binding of nicotine to the brain form of the nicotinic acetylcholine receptor compared to the peripheral form is produced by a point mutation outside of the binding pocket (Xiu et al., 2009). The mutation is thought to produce only slight alterations in geometry of the binding site that have major effects on ligand binding. Research results indicate that despite the accumulated toxicological facts in favor of a toxic gain of function for NTE in OPIDN, deleterious effects arise from a loss of function of the protein. For example, disruption of the Nte gene in mice resulting in effective knockout of NTE is lethal during gestation (Winrow et al., 2003), presumably from interference with vascular and placental development (Moser et al., 2004). Conditional knockout of brain NTE was not

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lethal but produced vacuolization of neuropil in CA1– CA3 regions of hippocampus and partial loss of cerebellar Purkinje cells; this study also demonstrated subcellular localization of NTE in the ER (Akassoglou et al., 2004). Likewise, Nte/ mice survived and had approximately half the NTE activity of wild-type mice, but they were noted to be hyperactive and to exhibit higher mortality than wild-type mice to the neuropathic OP compound ethyl n-octylphosphonofluoridate (EOPF). Wild-type mice given a sublethal dose of EOPF expected to inhibit NTE also exhibited hyperactivity; however, this is difficult to interpret, in part because Nte/ mice given the same dose of EOPF had reduced motor activity. Finally, three point mutations in human NTE have been linked to a form of motor neuron disease (Rainier et al., 2008). The functional correlate of these mutations is not known, but one of the mutations results in truncation of the protein at a point within the active site domain that would eliminate at least the esterase function. The other two mutations are each located within the active site domain of NTE and are considered to be tolerated substitutions. Preliminary experiments carried out in our laboratory by Nichole Hein using mutated NTE catalytic domain indicate that the other two point mutations result in a reduction of esterase activity (Hein, 2009). Taken together, although these molecular genetic studies of NTE do not shed direct light on its role in OPIDN, they point to pathological effects resulting from a loss of function of the protein. Although certain mutations in NTE are linked to a motor neuron disease that might arise from a loss of function of the enzyme, the possibility that mutated NTE is also toxic cannot be ruled out; the two possibilities are not mutually exclusive (Rainier et al., 2008). Indeed, a loss or a gain of function of NTE could each produce injurious effects by different mechanisms. For example, in Drosophila, mutation, deletion, and overexpression of the catalytically active NTE homologue, SWS, each result in neuropathology (Mühlig-Versen et al., 2005). Furthermore, despite the fact that the polyglutamine-repeat diseases, such as Huntington’s disease, have been considered to be the result of a toxic gain of function of a mutated protein, the exact mechanism accounting for the specific neurotoxicity of the protein has not been established. It has been shown that post-translational modifications, including phosphorylation, are involved in the toxic gain of function via alterations in the binding of protein partners to the nonmutated portion of the molecule (Lim et al., 2008; Pennuto et al., 2009). This work has led to the conclusion that specific neurodegeneration arises from a combination of a loss of function of the normal protein and a toxic gain of function of the altered protein. Such may turn out to be the case with NTE and OPIDN; however, given what has been learned from studies of polyglutamine repeat diseases, it will be critical to study the full-length protein, its post-translational modifications, and its interactions with binding partners to understand a possible gain-of-function mechanism.

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67.6  Role of neuropathy target esterase in organophosphorus ester-induced delayed neuropathy: a relative loss of function in phospholipid metabolism? Studies of NTE indicate that its physiological function entails its catalytic activity toward phospholipids, and that a loss of this function relative to that of other phospholipidmetabolizing enzymes might explain the pathogenesis of OPIDN (Glynn, 2006; Quistad et al., 2003; van Tienhoven et al., 2002; Zaccheo et al., 2004). Here, we expand upon these pioneering ideas to present what we call the “lipid hypothesis” or “lipid model” of OPIDN. In particular, it appears that NTE could be a lysophospholipase – a protein that hydrolyzes lysophospholipids (the lipid product of phospholipid hydrolysis by PLA2) to yield free fatty acids and a glycerophosphate (Quistad et al., 2003; van Tienhoven et al., 2002). NTE can act as a phospholipase as well as a lysophospholipase, with a preference toward the latter role, as shown by its hydrolysis of lysophosphatidylcholine (van Tienhoven et al., 2002). Lysophosphatidylcholine can also serve as a substrate for lysophospholipase D to produce lysophosphatidic acid (LPA), a molecule capable of a variety of biological effects, including mimicry of growth factors (Moolenaar, 2002; Moolenaar et al., 1997; Wang and Dennis, 1999) and blocking caspase-8 (Kang et al., 2004). In addition, in lysophospholipids such as LPA, the presence of a single hydrophobic fatty acid chain in these molecules enables them to act as detergents, which can disrupt membrane structure. Consequently, in healthy cells the membrane concentrations of lysophospholids are usually limited to 0.5–6% (by weight) of the total lipid content (Stafford et al., 1989; Wang and Dennis, 1999). Therefore, loss of homeostasis resulting in elevated lysophospholipid levels could jeopardize membrane structure and integrity through micelle formation and solubilization of membrane segments. If such disruption were to take place in the membrane of the ER, where NTE is localized (Glynn, 1999; Mühlig-Versen et al., 2005), it would be expected to lead to a loss of calcium homeostasis in the cell because the ER is the primary cellular calcium store (Verkhratsky, 2005). In fact, increased levels of intracellular calcium were reported following coadministration in vivo of the neuropathic OP compound phenyl saligenin phosphate along with the phenylalkylamine calcium channel blocker verapamil in adult hens (el-Fawal and Ehrich, 1993). Loss of calcium homeostasis in the cell could result in the unregulated activation of calpains, which would begin to break down the cytoskeleton and lead to accumulation of calcium in mitochondria, thereby initiating the formation of the mitochondrial permeability transition and apoptotic cell death. In the case of the neuron, an axonal self-destruct program appears to operate that has some analogies to apoptosis but that is distinct from it (Coleman, 2005;

Chapter | 67  Neuropathy Target Esterase

Raff et al., 2002). Given the apparent convergence of mechanisms of axonopathies initiated by a variety of insults, this proactive mechanism might be used to explain the axono­ pathy associated with OPIDN. As seen in Figure 67.5, the lipid hypothesis postulates that a neuropathic OP compound induces OPIDN via ­accumulation

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of lysophospholipids in the ER membrane, resulting in its disruption via the formation of lysophospholipid micelles, which would solubilize regions of the membrane. If this damage were of sufficient severity, it would be expected to result ultimately in axonopathy. The required levels of lysophospholipids could be achieved by striking the right balance of

Figure 67.5  Proposed mechanism for the lipid hypothesis. (A) The phospholipid pathway depicting the points at which an OP inhibitor can act. Hydrolysis of a phospholipid at the sn-2 position by PLA2 results in the formation of a lysophospholipid. In normal circumstances, the levels of lysophospholipids in the membrane are maintained at 0.5–6% (by weight) of the membrane’s total lipid content. This is accomplished by the proper balance of acyltransferases catalyzing the reacylation of lysophospholipids to phospholipids and lysophospholipases (e.g., NTE) catalyzing the hydrolysis of the remaining fatty acid ester bond resulting in the release of a fatty acid and the glycerophosphate head of the phospholipid. OP inhibitors can block PLA2, acyltransferases, and NTE (presumably along with other lysophospholipases). Thus, it is postulated that the differential rates of inhibition of these enzymes determine the neuropathic potential of an OP compound. (B) In the case of a nonneuropathic OP compound, it is postulated that the inhibitory potential of the OP compound against PLA2 is greater than that against NTE, other lysophospholipases (LysoPLA), or the acyltransferases. The relative inhibitory potencies of the OP compound against PLA2, acyltransferase, as well as NTE and other lysophospholipases are indicated by the length of the thick lines, with the longer lines indicating higher levels of inhibition. Thus, the rate of formation of lysophospholipids is lower than its clearance by acyltransferases and NTE. (C) On the other hand, a neuropathic OP compound is expected to be more potent against NTE/other lysophospholipases and/or the acyltransferases, relative to PLA2, thus leading to accumulation of lysophospholipids (LysoPL) in the membrane. The unregulated buildup of these lysophospholipids disrupts the stability of the ER membrane, thereby compromising its structure, which, if sufficiently severe, would be expected to lead to axonopathy. R, alkyl chain; X, polar head group (Wijeyesakere, 2009).

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i­nhibition of NTE versus other enzymes involved in metabol­ ism of these lipids. Thus, OPIDN would be precipitated by a relative loss of function of NTE rather than by a gain of function through formation of the aged protein. The lipid model predicts that neuropathic OP compounds should be more potent against NTE and/or the acyltransferases (enzymes that catalyze the acylation of lysophospholipids) than PLA2. In contrast, if a compound were more potent against PLA2, it would inhibit the formation of these lysophospholipids to a greater degree than their hydrolysis by NTE or their reacylation by acyltransferases. This prediction is supported, in part, by results indicating that neuropathic OP compounds such as chlorpyrifos oxon (CPO; the active metabolite of the insecticide chlorpyrifos, which is neuropathic at supralethal doses) and EOPF are more potent against NTE than PLA2. The reported I50 values (preincubation times not given) for CPO and EOPF against PLA2 are each in excess of 10,000 nM compared to values of 180 and 0.05 nM, respectively, for these compounds against NTE (Casida and Quistad, 2005; Casida et al., 2008). In addition, the lipid model predicts that nonneuropathic OP compounds should be more potent against PLA2 than NTE, other lysophospholipases, or the acyltransferases. Thus, treatment with nonneuropathic OP compounds would result in a slower rate of formation of lysophospholipids relative to their rate of clearance, taking into account the turnover rates for these enzymes. Given that lysophospholipids would not accumulate in the ER membrane, neuro­toxicity would be averted. Currently, however, data on relative inhibitory potencies of nonneuropathic compounds against the relevant enzymes are lacking. It follows from the lipid model that the relative inhibitory potentials of OP compounds against PLA2, NTE, other lysophospholipases, and the acyltransferases determine which OP compounds are neuropathic and which are not. To test the model, the lysophospholipids that serve as NTE substrates and the acyltransferases that reacylate them should be conclusively identified. It would also be important to identify lysophospholipases other than NTE that are able to catalyze the hydrolysis of the lysophospholipids that serve as physiological substrates for NTE. Alternatively, a metabolomic approach determining the concentrations of phospholipids and lysophospholipids at varying time intervals postexposure in cells treated with neuropathic and nonneuropathic OP compounds could be used to investigate the validity of the proposed lipid hypothesis. Such research has important regulatory implications because if the lipid hypothesis for OPIDN were shown to be correct, the current U.S. EPA guidelines for testing the neuropathic potential of OP compounds, which currently specify testing NTE inhibition ex vivo (U.S. EPA, 1991), would need to be revised to include testing inhibitory potentials against PLA2, important lysophospholipases other than NTE, and the relevant acyltransferases. Although the ability of an NTE inhibitor to undergo aging after it has formed the NTE–inhibitor conjugate is currently

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regarded as a determinant of its neuropathic potential, the lipid model indicates that aging per se is not the causative event in the initiation of OPIDN. Therefore, there might be nonaging NTE inhibitors that could inhibit NTE to a greater degree than PLA2, thereby leading to OPIDN. Although this area has not been researched extensively, as noted previously, there have been published clinical case reports of delayed neurotoxicity allegedly arising from the ingestion of carbamates (Dickoff et al., 1987; Umehara et al., 1991). However, one must be cautious and not attribute too much significance to these reports because they are not the result of controlled experiments and could simply represent cases in which the ingested carbamates were contaminated with conventional neuropathic OP compounds. This represents an area in which further research is needed before a conclusive statement can be made regarding the validity of these case findings. A potential complication faced by the proposed lipid hypothesis concerns the issue of enzyme turnover, particularly the turnover of NTE. Any model for OPIDN has to explain the postexposure 1- to 4-week asymptomatic period in humans, which is characteristic of this condition. In our hands, in experimental studies on hens, the average delay between dosing and onset of unequivocal clinical signs is approximately 11 days. The half-life for the turnover of NTE in hens has been reported to be approximately 5 days in the brain and 7 days in the spinal cord (Meredith and Johnson, 1988). Thus, based on classical pharmacokinetic principles, turnover would be essentially complete within four to seven half-lives so that inhibited NTE should be entirely replaced within 20–28 days in brain and 28–49 days in spinal cord. Depending on the level of inhibition required to maintain accumulation of lysophospholipids, these times would fall within the duration of the observed delay between exposure and appearance of clinical signs. Although the issue of enzyme turnover may present a challenge to the acceptance of the lipid hypothesis, the same issue exists with the toxic gain-of-function model based on aged NTE. Thus, if a persistent insult from the aged enzyme is needed during the development of toxicity, replacement of aged NTE would be expected to halt the insult to the neuron, thereby allowing it to recover. However, turnover of NTE has been measured in terms of return of enzyme activity after inhibition rather than the presence of aged protein. Therefore, it would be of interest to determine if the aged protein were cleared at a slower rate than normal protein, thus allowing the putative toxic form of NTE to remain longer in the cell to exert an injurious influence. Both the lipid model and the aged NTE model of OPIDN could get around the question of the delay between exposure and appearance of clinical signs by invoking a triggering mechanism such as an irreversible signaling cascade (Coleman, 2005; Raff et al., 2002). In the case of the lipid hypothesis, once a critical concentration of damaging lysophospholipids were reached in the ER, the triggering event might be ER stress, as has been observed in certain

Chapter | 67  Neuropathy Target Esterase

neurodegenerative diseases (Ito and Suzuki, 2009) and produced by a lysophospholipid analog in cancer cells (NietoMiguel et al., 2007).

Conclusion Human NTE is now known as patatin-like phospholipase domain-containing protein-6 (PLPL6_HUMAN) because of the homology of its catalytic domain to patatin, a lipid acyl hydrolase found in potatoes and other plants. It is also cataloged as a lysophospholipase (EC 3.1.1.5) due to its hydrolysis of lysophosphatidylcholine to glycerophosphocholine, but its precise physiological and pathogenic roles have not been firmly established. The originally reported sequence is currently named isoform-2 (a splice variant); it contains 1327 amino acid residues, bringing the molecular weight with no post-translational modification to 146 kDa. NTE has a transmembrane domain near the N-terminus, three putative cyclic nucleotide-binding domains, and the patatin esterase domain near the C-terminus. Homology modeling indicates that the active site consists of a serine–aspartate dyad rather than a classical catalytic triad. The protein was first identified in hen brain as the target for initiation of OPIDN by neuropathic OP compounds. NTE is commonly assayed as the activity toward the nonphysiological substrate PV that is resistant to preincubation with nonneuropathic paraoxon and sensitive to neuropathic mipafox. Extensive structure–activity relationships demonstrated that NTE inhibitors fell into two functional categories: neuropathic and nonneuropathic. Neuropathic inhibitors included certain phosphates, phosphinates, and phosphoramidates, whereas nonneuropathic inhibitors included particular phosphinates, carbamates, and sulfonyl fluorides. The essential difference between neuropathic and nonneuro­ pathic NTE inhibitors is that neuropathic inhibitors are capable of undergoing the aging reaction, whereby the NTE–OP conjugate undergoes net dealkylation/dearylation (phosphates or phosphonates) or deprotonation (phosphoramidates) to yield a negatively charged phosphyl group covalently bonded to the active site serine. Nonneuropathic NTE inhibitors are incapable of aging, but they are far from being biologically inert. Pretreatment of test animals with these nonneuropathic NTE inhibitor compounds protects against subsequent administration of neuropathic NTE inhibitors, presumably by blocking the formation of aged enzyme. Neuropathic potential of OP compounds can be assessed by dosing adult hens and assaying brain NTE inhibition 24 h later ex vivo. A threshold level of approximately 70% inhibition must be exceeded for clinical signs of OPIDN to be apparent 1–4 weeks later; the average time to unequivocal clinical onset in hens is approximately 11 days. Repeated dosing results in NTE inhibition reaching a plateau; if the steady-state level is below threshold, clinical OPIDN will not ensue. Predictions of neuropathic potential can also be made

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in vitro by determining RIPs of the directly active form of the OP compound against AChE and NTE. Potential for aging can be assessed by inspection of the chemical structure of the compound or by biochemical testing for reactivatability by powerful nucleophiles, such as KF. Although serum, plasma, and erythrocytes lack NTE, lymphocytes possess NTE activity that has been used as a biomarker of exposure to potentially neuropathic OP compounds. Because initiation of OPIDN appears to require aging of NTE rather than mere inhibition of its activity, it has been thought that aged NTE operates through a toxic gainof-function mechanism, possibly by acting as a spurious phosphorylation signal to trigger a proactive axonopathic process akin to Wallerian degeneration. Conventional and conditional knockout experiments have shown NTE to be essential for the development and maintenance of various cells and tissues, including large neurons in the hippocampus and cerebellum. In addition, although Nte/ mice with approximately half the normal levels of NTE survive and appear to undergo normal morphological development of the nervous system, they exhibit hyperactivity and increased sensitivity to the lethal effects of a neuropathic OP compound, EOPF. Moreover, mutations in the catalytic domain of human NTE have been linked to a form of motor neuron disease; one of the mutations results in truncation of the protein, and preliminary studies indicate that the other two mutations produce decreased enzymatic activity. Taken together, the molecular genetic results suggest that a loss of function of NTE could be injurious. However, the NTE mutation results do not rule out a possible toxic gain of function of the mutated protein. The lipid model of OPIDN suggests that NTE acts in concert with other lysophospholipases, PLA2, and acyltransferases to regulate levels of toxic lysophospholipids in ER membranes. In this paradigm, the relative potency of an OP compound against these enzymes would determine if the compound were neuropathic or not. Although the model provides an alternative to the toxic gain-of-function hypothesis of aged NTE, further research is needed to decide between the options or to determine that they are not mutually exclusive and that both mechanisms may be operating.

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Johnson, M. K. (1978). The anomalous behaviour of some dimethylphosphates in the biochemical test for delayed neurotoxicity potential. Arch. Toxicol. 41, 107–110. Johnson, M. K. (1981). Delayed neurotoxicity – do trichlorphon and/or dichlorvos cause delayed neuropathy in man or in test animals?  Acta Pharmacol. Toxicol. 49, 87–98. Johnson, M. K. (1982a). The target for initiation of delayed neurotoxicity by organophosphorus esters: biochemical studies and toxicological applications. In “Reviews in Biochemical Toxicology” (E. Hodgson, J. R. Bend, and R. M. Philpot, eds.), Vol. 4, pp. 141–212. Elsevier, New York. Johnson, M. K. (1982b). Check your paraoxon and parathion for neurotoxic impurities. Vet. Hum. Toxicol. 24, 220a. Johnson, M. K. (1984). Delayed neurotoxicity tests of organophosphorus esters: a proposed protocol integrating neuropathy target esterase (NTE) assays with behaviour and histopathology tests to obtain more information more quickly from fewer animals. In “Proceedings of the International Conference on Environmental Hazards of Agrochemicals in Developing Countries, Alexandria, Egypt, November 8–12, 1983” (A. H. El-Sebaed, ed.), pp. 474–493. University of Alexandria, Alexandria, Egypt. Johnson, M. K. (1987). The importance of chirality in influencing both acute and delayed neuropathic toxicity of organophosphorus esters. Toxicol. Environ. Chem. 14, 321–335. Johnson, M. K. (1988). Sensitivity and selectivity of compounds interacting with neuropathy target esterase: further structure/activity studies. Biochem. Pharmacol. 37, 4095–4104. Johnson, M. K. (1990). Organophosphates and delayed neuropathy – Is NTE alive and well? Toxicol. Appl. Pharmacol. 102, 385–399. Johnson, M. K., and Lauwerys, R. R. (1969). Protection by some carbamates against the delayed neurotoxic effect of di-isopropyl phosphorofluoridate. Nature (London) 222, 1066–1067. Johnson, M. K., and Lotti, M. (1980). Delayed neurotoxicity caused by chronic feeding of organophosphates requires a high-point of inhibition of neurotoxic esterase. Toxicol. Lett. 5, 99–102. Johnson, M. K., and Read, D. J. (1987). The influence of chirality on the delayed neuropathic potential of some organophosphorus esters: neuropathic and prophylactic effects of stereoisomers of ethyl phenylphosphonic acid (EPN oxon and EPN) correlate with quantities of aged and unaged neuropathy target esterase in vivo. Toxicol. Appl. Pharmacol. 90, 103–115. Johnson, M. K., Vilanova, E., and Read, D. J. (1991). Anomalous biochemical responses in tests of the delayed neuropathic potential of methamidophos (O, S-dimethyl phosphorothioamidate), its resolved isomers and of some higher O-alkyl homologues. Arch. Toxicol. 65, 618–624. Kang, Y. C., Kim, K. M., Lee, K. S., Namkoong, S., Lee, S. J., Han, J. A., Jeoung, D., Ha, K. S., Kwon, Y. G., and Kim, Y. M. (2004). Serum bioactive lysophospholipids prevent TRAIL-induced apoptosis via PI3K/Akt-dependent cFLIP expression and Bad phosphorylation. Cell Death Differ. 11, 1287–1298. Kayyali, U. S., Moore, T. B., Randall, J. C., and Richardson, R. J. (1991). Neurotoxic esterase (NTE) assay: optimized conditions based on detergent-induced shifts in the phenol/4-aminoantipyrine chromophore spectrum. J. Anal. Toxicol. 15, 86–89. Kohli, N., Srivastava, D., Sun, J., Richardson, R. J., Lee, I. S., and Worden, R. M. (2007). Nanostructured biosensor for measuring neuro­ pathy target esterase activity. Anal. Chem. 79, 5196–5203. Kretzschmar, D., Hasan, G., Sharma, S., Heisenberg, M., and Benzer, S. (1997). The Swiss cheese mutant causes glial hyperwrapping of and brain degeneration in Drosophila. J. Neurosci. 17, 7425–7432.

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Kropp, T. J., and Richardson, R. J. (2003). Relative inhibitory potencies of chlorpyrifos oxon, chlorpyrifos methyl oxon, and mipafox for acetylcholinesterase versus neuropathy target esterase. J. Toxicol. Environ. Health A 66, 1145–1157. Kropp, T. J., Glynn, P., and Richardson, R. J. (2004). The mipafox-inhibited catalytic domain of human neuropathy target esterase ages by reversible proton loss. Biochemistry 43, 3716–3722. Lim, J., Crespo-Barreto, J., Jafar-Nejad, P., Bowman, A. B., Richman, R., Hill, D. E., Orr, H. T., and Zoghbi, H. Y. (2008). Opposing effects of polyglutamine expansion on native protein complexes contribute to SCA1. Nature 452, 713–718. Lotti, M. (1992). The pathogenesis of organophosphate polyneuropathy. Crit. Rev. Toxicol. 21, 465–487. Lotti, M. (1995). Cholinesterase inhibition: complexities in interpretation. Clin. Chem. 41, 1814–1818. Lotti, M., and Johnson, M. K. (1978). Neurotoxicity of organophosphorus pesticides: Predictions can be based on in vitro studies with hen and human enzymes. Arch. Toxicol. 41, 215–221. Lotti, M., and Johnson, M. K. (1980a). Repeated small doses of a neuro­ toxic organophosphate: Monitoring of neurotoxic esterase in brain and spinal cord. Arch. Toxicol. 45, 263–271. Lotti, M., and Johnson, M. K. (1980b). Neurotoxic esterase in human nervous tissue. J. Neurochem. 34, 747–749. Lotti, M., and Moretto, A. (1999). Promotion of organophosphate induced polyneuropathy by certain esterase inhibitors. Chem. Biol. Interact. 119120, 519–524. Lotti, M., Ferrara, S. D., Caroldi, S., and Sinigaglia, F. (1981). Enzyme studies with human and hen autopsy tissue suggest omethoate does not cause delayed neuropathy in man. Arch. Toxicol. 48, 265–270. Lotti, M., Becker, C. E., Aminoff, M. J., Woodrow, J. E., Seiber, J. N., Talcott, R. E., and Richardson, R. J. (1983). Occupational exposure to the cotton defoliants DEF and merphos. A rational approach to monitoring organophosphorus-induced delayed neurotoxicity. J. Occup. Med. 25, 517–522. Lotti, M., Moretto, A., Zoppellari, R., Dainese, R., Rizzuto, N., and Barusco, G. (1986). Inhibition of lymphocytic neuropathy target esterase predicts the development of organophosphate-induced delayed polyneuropathy. Arch. Toxicol. 59, 176–179. Lush, M. J., Li, Y., Read, D. J., Willis, A. C., and Glynn, P. (1998). Neuropathy target esterase and a homologous Drosophila neuro­ degeneration mutant protein contain a domain conserved from bacteria to man. Biochem. J. 332, 1–4. Mackay, C. E., Hammock, B. D., and Wilson, B. W. (1996). Identification and isolation of a 155 kDa protein with neuropathy target esterase activity. Fundam. Appl. Toxicol. 30, 23–30. Makhaeva, G. F., Sigolaeva, L. V., Zhuravleva, L. V., Eremenko, A. V., Kuorchkin, I. N., Richardson, R. J., and Malygin, V. V. (2003). Biosensor detection of neuropathy target esterase in whole blood as a biomarker of exposure to neuropathic organophosphorus compounds. J. Toxicol. Environ. Health A 66, 599–610. Makhaeva, G. F., Malygin, V. V., Strakhova, N. N., Sigolaeva, L. V., Sokolovskaya, L. G., Eremenko, A. V., Kurochkin, I. N., and Richardson, R. J. (2007). Biosensor assay of neuropathy target esterase in whole blood as a new approach to OPIDN risk assessment: review of progress. Hum. Exp. Toxicol. 26, 273–282. Maroni, M., and Bleecker, M. L. (1986). Neuropathy target esterase in human lymphocytes and platelets. J. Appl. Toxicol. 6, 1–7. McConnell, R., Delgado-Téllez, E., Cuadra, R., Tórres, E., Keifer, M., Almendárez, J., Miranda, J., El-Fawal, H. A., Wolff, M., Simpson, D., and Lundberg, I. (1999). Organophosphate neuropathy due to

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­ ethamidophos: biochemical and neurophysiological markers. Arch. m Toxicol. 73, 296–300. Meredith, C., and Johnson, M. K. (1988). Neuropathy target esterase: Rates of turnover in vivo following covalent inhibition with phenyl di-n-pentylphosphinate. J. Neurochem. 51, 1097–1101. Milatovic, D., Moretto, A., Osman, K. A., and Lotti, M. (1997). Phenyl valerate esterases other than neuropathy target esterase and the promotion of organophosphate polyneuropathy. Chem. Res. Toxicol. 10, 1045–1048. Moolenaar, W. H. (2002). Lysophospholipids in the limelight: autotaxin takes center stage. J. Cell Biol. 158, 197–199. Moolenaar, W. H., Kranenburg, O., Postma, F. R., and Zondag, G. C. (1997). Lysophosphatidic acid: G-protein signalling and cellular responses. Curr. Opin. Cell Biol. 9, 168–173. Moretto, A., and Lotti, M. (1998). Poisoning by organophosphorus insecticides and sensory neuropathy. J. Neurol. Neurosurg. Psychiatr. 64, 463–468. Moretto, A., Nicolli, A., and Lotti, M. (2005). Peripheral nerve esterases and the promotion of organophosphate-induced neuropathy in hens. Chem. Biol. Interact. 157158, 285–291. Moretto, A., Nicolli, A., and Lotti, M. (2007). The search of the target of promotion: phenylbenzoate esterase activities in hen peripheral nerve. Toxicol. Appl. Pharmacol. 219, 196–201. Moser, M., Stempl, T., Li, Y., Glynn, P., Buttner, R., and Kretzschmar, D. (2000). Molecular cloning of the mouse NTE/SWS gene. Mech. Dev. 90, 279–282. Mühlig-Versen, M., da Cruz, A. B., Tschäpe, J. A., Moser, M., Büttner, R., Athenstaedt, K., Glynn, P., and Kretzschmar, D. (2005). Loss of Swiss cheese/neuropathy target esterase activity causes disruption of phosphatidyl­choline homeostasis and neuronal and glial death in adult Drosophila. J. Neurosci. 25, 2865–2873. Nardini, M., and Dijkstra, B. W. (1999). a/ß Hydrolase fold enzymes: the family keeps growing. Curr. Opin. Struct. Biol. 9, 732–737. Nieto-Miguel, T., Fonteriz, R. I., Vay, L., Gajate, C., López-Hernández, S., and Mollinedo, F. (2007). Endoplasmic reticulum stress in the proapoptotic action of edelfosine in solid tumor cells. Cancer Res. 67, 10368–10378. Novak, R., and Padilla, S. (1986). An in vitro comparison of rat and chicken brain neurotoxic esterase. Fundam. Appl. Toxicol. 6, 464–471. Organization for Economic Co-operation and Development (OECD) (1995). Delayed neurotoxicity of organophosphorus substances following acute exposure and Delayed neurotoxicity of organophosphorus substances: 28-day repeated dose study. In “Guidelines for Testing of Chemicals 418 and 419.” Environmental Health and Safety Division, OECD, Paris. Pennuto, M., Palazzolo, I., and Poletti, A. (2009). Post-translational modifications of expanded polyglutamine proteins: impact on neurotoxi­ city. Hum. Mol. Genet. 18, R40–R47. Pope, C. N., and Padilla, S. S. (1989a). Modulation of neurotoxic esterase activity in vitro by phospholipids. Toxicol. Appl. Pharmacol. 97, 272–278. Pope, C. N., and Padilla, S. S. (1989b). Chromatographic characterization of neurotoxic esterase. Biochem. Pharmacol. 38, 181–188. Poulsen, E., and Aldridge, W. N. (1964). Studies on esterases in the chicken nervous system. Biochem. J. 90, 182–189. Quistad, G. B., and Casida, J. E. (2004). Lysophospholipase inhibition by organophosphorus toxicants. Toxicol. Appl. Pharmacol. 196, 319–326. Quistad, G. B., Barlow, C., Winrow, C. J., Sparks, S. E., and Casida, J. E. (2003). Evidence that mouse brain neuropathy target esterase is a lysophospholipase. Proc. Natl. Acad. Sci. USA 100, 7983–7987. Raff, M. C., Whitmore, A. V., and Finn, J. T. (2002). Axonal self-destruction and neurodegeneration. Science 296, 868–871.

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Rainier, S., Bui, M., Mark, E., Thomas, D., Tokarz, D., Ming, L., Delaney, C., Richardson, R. J., Albers, J. W., Matsunami, N., Stevens, J., Coon, H., Leppert, M., and Fink, J. K. (2008). Neuropathy target esterase gene mutations cause motor neuron disease. Am. J. Hum. Genet. 82, 780–785. Randall, J. C., Yano, B. L., and Richardson, R. J. (1997). Potentiation of organophosphorus compound-induced delayed neurotoxicity (OPIDN) in the central and peripheral nervous system of the adult hen: Distribution of axonal lesions. J. Toxicol. Environ. Health 51, 571–590. Richardson, R. J. (1984). Neurotoxic esterase: normal and pathogenic roles. In “Cellular and Molecular Neurotoxicology” (T. Narahashi, ed.), pp. 285–295. Raven Press, New York. Richardson, R. J. (1992). Interactions of organophosphorus compounds with neurotoxic esterase. In “Organophosphates: Chemistry, Fate and Effects” (J. E. Chambers and P. E. Levi, eds.), pp. 299–323. Academic Press, San Diego. Richardson, R. J. (1995). Assessment of the neurotoxic potential of chlorpyrifos relative to other organophosphorus compounds: a critical review of the literature. J. Toxicol. Environ. Health 44, 135–165. Richardson, R. J. (2005). Neurotoxicity, delayed. In “Encyclopedia of Toxicology” (P. Wexler, ed.), 2nd ed., Vol. 2, pp. 302–306. Academic Press, New York. Richardson, R. J., and Dudek, B. R. (1983). Neurotoxic esterase: characterization and potential for a predictive screen for exposure to neuro­ pathic organophosphates. In “Pesticide Chemistry: Human Welfare and the Environment” (J. Miyamoto and P. C. Kearney, eds.), Vol. 3, pp. 491–495. Pergamon, Oxford. Richardson, R. J., Davis, C. S., and Johnson, M. K. (1979). Subcellular distribution of marker enzymes and of neurotoxic esterase in adult hen brain. J. Neurochem. 32, 607–615. Richardson, R. J., Moore, T. B., Kayyali, U. S., Fowke, J. H., and Randall, J. C. (1993a). Inhibition of hen brain acetylcholinesterase and neuro­ toxic esterase by chlorpyrifos in vivo and kinetics of inhibition by chlorpyrifos oxon in vitro: application to assessment of neuropathic risk. Fundam. Appl. Toxicol. 20, 273–279. Richardson, R. J., Moore, T. B., Kayyali, U. S., and Randall, J. C. (1993b). Chlorpyrifos: assessment of potential for delayed neurotoxicity by repeated dosing in adult hens with monitoring of brain acetylcholin­ esterase, brain and lymphocyte neurotoxic esterase, and plasma butyryl­cholinesterase activities. Fundam. Appl. Toxicol. 21, 89–96. Richardson, R. J., Worden, R. M., and Makhaeva, G. F. (2009). Biomarkers and biosensors of delayed neuropathic agents. In “Handbook of Toxicology of Chemical Warfare Agents” (R. C. Gupta, ed.), pp. 859–876. Elsevier, Amsterdam. Rueffer-Turner, M. E., Read, D. J., and Johnson, M. K. (1992). Purification of neuropathy target esterase from avian brain after pre­ labelling with [3H]di-isopropyl phosphorofluoridate. J. Neurochem. 58, 135–141. Rydel, T. J., Williams, J. M., Krieger, E., Moshiri, F., Stallings, W. C., Brown, S. M., Pershing, J. C., Purcell, J. P., and Alibhai, M. F. (2003). The crystal structure, mutagenesis, and activity studies reveal that patatin is a lipid acyl hydrolase with a Ser-Asp catalytic dyad. Biochemistry 42, 6696–6708. Schwab, B. W., and Richardson, R. J. (1986). Lymphocyte and brain neurotoxic esterase: dose and time dependence of inhibition in the hen examined with three organophosphorus esters. Toxicol. Appl. Pharmacol. 83, 1–9. Senanayake, N., and Johnson, M. K. (1982). Acute polyneuropathy after poisoning by a new organophosphorus insecticide. N. Engl. J. Med. 306, 155–156.

Chapter | 67  Neuropathy Target Esterase

Smith, M. I., Elvove, E., and Frazier, W. H. (1930). The pharmacological action of certain phenol esters with special reference to the etiology of the so-called ginger paralysis. Public Health Rep. 45, 2509–2524. Sogorb, M. A., Viniegra, S., Reig, J. A., and Vilanova, E. (1994). Partial characterization of neuropathy target esterase and related phenyl valerate esterases from bovine adrenal medulla. J. Biochem. Toxicol. 9, 145–152. Sogorb, M. A., Bas, S., Gutierrez, L. M., Vilanova, E., and Viniegra, S. (1997). Bovine chromaffin cells as an in vitro model for the study of non-cholinergic toxic effects of organophosphorus compounds. Arch. Toxicol. Suppl. 19, 347–355. Soliman, S. A., Curley, A., Farmer, J., and Novak, R. (1986). In vivo inhibition of chicken brain acetylcholinesterase and neurotoxic esterase in relation to the delayed neurotoxicity of leptophos and cyanophenphos. J. Environ. Pathol. Toxicol. Oncol. 7, 211–224. Stafford, R. E., Fanni, T., and Dennis, E. A. (1989). Interfacial properties and critical micelle concentration of lysophospholipids. Biochemistry 28, 5113–5120. Taylor, P. (1992). Impact of recombinant DNA technology and protein structure determination on past and future studies on acetylcholinesterase. In “Multi-disciplinary Approaches to Cholinesterase Functions” (A. Shafferman and B. Velan, eds.), pp. 1–15. Plenum, New York. Thomas, T. C., Szekacs, A., Rojas, S., Hammock, B. D., Wilson, B. W., and MacNamee, M. G. (1990). Characterization of neuropathy target esterase using trifluoromethyl ketones. Biochem. Pharmacol. 40, 2587–2596. Thompson, C. M., and Richardson, R. J. (2004). Anticholinesterase insecticides. In “Pesticide Toxicology and International Regulation” (T. C. Marrs and B. Ballantyne, eds.), pp. 89–127. Wiley, Chichester, UK. Umehara, F., Izumo, S., Arimura, K., and Osame, M. (1991). Polyneuropathy induced by m-tolyl methyl carbamate intoxication. J. Neurol. 238, 47–48. U.S. Environmental Protection Agency (EPA) (1991). “Pesticide Assessment Guidelines, Subdivision F; Hazard Evaluation: Human and Domestic Animals: Addendum 10, Neurotoxicity,” Series 81–83, pp. 3-12, EPA 540/09-91-123, PB 91-154617. Health Effects Division, Office of Pesticide Programs, U.S. EPA, Washington, DC. van Tienhoven, M., Atkins, J., Li, Y., and Glynn, P. (2002). Human neuro­pathy target esterase catalyzes hydrolysis of membrane lipids. J. Biol. Chem. 277, 20942–20948. Verkhratsky, A. (2005). Physiology and pathophysiology of the calcium store in the endoplasmic reticulum of neurons. Physiol. Rev. 85, 201–279. Veronesi, B., Padilla, S., and Lyerly, D. (1986). The correlation between neurotoxic esterase inhibition and mipafox-induced neuropathic damage in rats. Neurotoxicology 7, 207–215. Veronesi, B., Padilla, S., Blackmon, K., and Pope, C. (1991). Murine susceptibility to organophosphorus-induced delayed neuropathy (OPIDN). Toxicol. Appl. Pharmacol. 107, 311–324.

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Veronesi, B., Ehrich, M., Blusztajn, J. K., Oortgiesen, M., and Durham, H. (1997). Cell culture models of interspecies selectivity to organophosphorus insecticides. Neurotoxicology 18, 283–297. Vilanova, E., Escudero, M. A., and Barril, J. (1999). NTE soluble isoforms: new perspectives for targets of neuropathy inducers and promoters. Chem. Biol. Interact. 119120, 525–540. Vose, S. C., Fujioka, K., Gulevich, A. G., Lin, A. Y., Holland, N. T., and Casida, J. E. (2008). Cellular function of neuropathy target esterase in lysophosphatidylcholine action. Toxicol. Appl. Pharmacol. 232, 376–383. Wang, A., and Dennis, E. A. (1999). Mammalian lysophospholipases. Biochim. Biophys. Acta 1439, 1–16. Wijeyesakere, S. J. (2009). Enzyme aging: structural insights from NTE and patatin. PhD dissertation, University of Michigan, Ann Arbor. Wijeyesakere, S. J., Richardson, R. J., and Stuckey, J. A. (2007a). Modeling the tertiary structure of the patatin domain of neuropathy target esterase. Protein J. 26, 165–172. Wijeyesakere, S. J., Stuckey, J. A., and Richardson, R. J. (2007b). Molecular modeling predicts conformational changes in the neuro­ pathy target esterase (NTE) catalytic domain (PNTE) after inhibition and aging [abstract]. Toxicologist 96, 186. Wijeyesakere, S. J., Stuckey, J. A., and Richardson, R. J. (2009). Structural insight into inhibition/aging of neuropathy target esterase (NTE) from x-ray crystal studies of its catalytic domain homologue, patatin-17 (pat17) [abstract]. Toxicologist 108, 286–287. Williams, D. G., and Johnson, M. K. (1981). Gel electrophoretic identification of hen brain neurotoxic esterase labelled with tritiated di-isopropyl phosphorofluoridate. Biochem. J. 199, 323–333. Wilson, B. W., and Henderson, J. D. (1992). Blood esterase determinations as markers of exposure. Rev. Environ. Contam. Toxicol. 128, 55–69. Winrow, C. J., Hemming, M. L., Allen, D. M., Quistad, G. B., Casida, J. E., and Barlow, C. (2003). Loss of neuropathy target esterase in mice links organophosphate exposure to hyperactivity. Nat. Genet. 33, 477–485. Xiu, X., Puskar, N. L., Shanata, J. A. P., Lester, H. A., and Dougherty, D. A. (2009). Nicotine binding to brain receptors requires a strong cation- interaction. Nature 458, 534–537. Yoshida, M., Tomizawa, M., Wu, S.-Y., Quistad, G. B., and Casida, J. E. (1995). Neuropathy target esterase of hen brain: Active site reactions with 2-[octyl-3H]octyl-4H-1,3,2-benzodioxaphosphorin 2oxide and 2-octyl-4H-1,3,2-[aryl-3H]benzodioxaphosphorin 2-oxide. J. Neurochem. 64, 1680–1687. Zaccheo, O., Dinsdale, D., Meacock, P. A., and Glynn, P. (2004). Neuropathy target esterase and its yeast homologue degrade phosphatidylcholine to glycerophosphocholine in living cells. J. Biol. Chem. 279, 24024–24033.

Chapter 68

Cholinesterases Barry W. Wilson University of California, Davis, California

68.1  Introduction Cholinesterases (ChEs) are specialized carboxylic ester hydrolases that break down esters of choline. Two of spe­ cial concern to the pesticide toxicologist are acetylcholin­ esterase (AChE; acetylcholine hydrolase, EC 3.1.1.7) and butyrylcholinesterase (BuChE; acylcholine acylhydro­ lase, EC 3.1.1.8), also known as nonspecific cholinester­ ase or pseudocholinesterase. The preferred substrate for AChE is acetylcholine (ACh). Nonspecific cholinesterases prefer butyrylcholine and/or propionylcholine, depend­ ing on the species (Silver, 1974). This chapter discusses these enzymes, their importance in understanding the toxi­ city of organophosphate ester (OP) and carbamate (CB) pesticides, and their application to risk assessment (Taylor, 1999). ChEs are classed among the B-esterases, enzymes inhibited by OPs and possessing a serine catalytic site (Aldridge and Reiner, 1972; Ballantyne and Marrs, 1992; Chambers and Levi, 1992; Ecobichon, 1996; Gallo and Lawryk, 1991). Other B-esterases include the carboxyl­ esterases (CarbE, EC 3.1.1.1.), one of which is neuropathy target esterase (NTE), the enzyme associated with organo­ phosphate-induced delayed neuropathy (OPIDN) discussed in other chapters. The A-esterases are a different group of enzymes (e.g., arylesterases, paraoxonases, and DFPases) that actively hydrolyze OPs, providing an important means of detoxification (Furlong et al., 2000; Haley et al., 1999; La Du et al., 1999). There has been a great amount of research on ChEs since 1914 when Sir Henry Dale (Dale, 1914) proposed an esterase capable of hydrolyzing ACh in blood, and when Abderhalden and Paffrath (1925) and Loewi and Navratil (1926) prepared tissue extracts that broke down the chemical. In the past 30 years, the tertiary structure, amino acid, and DNA sequences of several ChEs have been elucidated. From 2000 to the present, approximately 14,000 research reports on ChEs have been listed in SciFinder Scholar, an ACS online database. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

Today, techniques such as site-directed mutagenesis, knockout mutants, and high-resolution x-ray crystallogra­ phy enable investigators to dissect the form and function of these proteins literally one amino acid at a time (Faerman et al., 1996; Gnatt et al., 1994; Silman and Sussman, 2008). In the future, this knowledge will help researchers design chemicals specifically targeted for the tertiary struc­ ture of these proteins and their genes. Specific reviews and conferences (Doctor et al., 1998; Massoulie et al., 1999; Reiner et al., 1999; Silman and Sussman, 2008; Taylor, 1994, 1996) and even full-color molecular structures dis­ played on the Internet help bring the reader the latest information from this rapidly growing research area. [For brevity, only selected references to a topic are cited here. The reader is referred to these references and to articles (e.g., Gallo and Lawryk, 1991) in previous editions of this book for citations to earlier work.] OPs with high toxicity were developed as chemical war­ fare agents in the late 1930s and early 1940s (Ecobichon, 1996; Holmstedt, 1963; Koelle, 1963). Since the 1950s, their offspring have been adapted as pesticides for agri­ cultural use (Ecobichon, 1996). Because of their potential as weapons, much research has focused on antidotes (e.g., oximes) and prophylactics to OP chemical warfare agents (National Academy of Sciences, 1999; Romano et al., 2008).

68.2  Distribution ChEs are widely distributed across animal species (Ecobichon, 1996). Their presence in insects and other inver­ tebrates has made anti-ChE agents popular and effective pesticides. Molecular forms of ChE similar to those in verte­ brates have been studied in animals as varied as nematodes (e.g., Caenorhabditis; Culetto et al., 1999), squid (Talesa et al., 1999), and Amphioxus, a protochordate (Pezzementi et al., 1998). A vertebrate-like AChE form has been reported in Paramecium, a ciliated protozoan (Corrado et al., 1999, 2005). 1457

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AChEs in the nervous system regulate excitation by destroying the neurotransmitter ACh. They are found at synapses and neuromuscular and myotendinous junctions and in cerebrospinal fluid, the central nervous system (CNS) neuron cell bodies, axons, and skeletal and smooth muscles (Silver, 1974). AChEs also are present in erythro­ cytes [red blood cells (RBCs)] of mammals, megakaryo­ cytes, lymphocytes, and platelets (Husain, 1994; Paulus et al., 1981; Zajicek, 1957). Although, there has been some interest in ChE in saliva (Ryhanen et al., 1983; Yamalik et al., 1990), the activity reported is low (Henn et al., 2006). Blood ChE forms are often used as surrogates for CNS ChEs in studies of toxicants. The AChE activity of human blood is restricted to its formed elements; most of it is vested in the RBCs (Wills, 1972). Plasma ChEs of other vertebrates often also hydrolyze ACh (Augustinsson, 1948, 1959a,b), and specific AChE enzyme activity is present in the plasma of some mammals. For example, the plasma ChE activity of rodents such as the laboratory rat is high in both AChE and BuChE (Traina and Serpietri, 1984). Neglecting this plasma AChE activity, as is often done, may lead to misinterpreting the extent of ChE inhibition in animals used for pesticide research and in the setting of regulations for food safety and human exposure (Wilson et al., 1996). AChE activity also has been found in the serum of mammalian embryos and birds; it decreases to adult levels after birth. AChE activity in fetal calf serum is high enough to be a source for purifying the enzyme (De la Hoz et al., 1986). In contrast, adult bovine blood has rel­ atively high RBC AChE and very low plasma ChE levels (Zajicek, 1957). Other species have a mixture of ChEs, first established by studies of substrate and inhibitor specifici­ ties and, recently, by DNA analyses (Bartels et al., 2000). Table 68.1 compares adult levels of RBC AChE from several species. The human has the highest activity. The rat has one of the lowest RBC AChE levels even though it is often the species used in biomedical research on anti-ChE agents. BuChEs and AChEs are also found together at syn­ apses, motor endplates, and muscle fibers (Silver, 1974). BuChE activity in blood is restricted to serum. The physiological functions of RBC and serum ChEs are unclear. The primary structure of ChEs is homologous to proteases and lipases (Taylor, 1996). One possibility is that blood ChEs evolved to protect the body from natural anti-ChE agents. A number of plant toxins have anti-ChE activity, including the solanaceous glycoalkaloids, which are naturally occurring steroids in potatoes and related plants (Krasowski et al., 1997; McGehee et al., 2000), and the fungal territrems (Chen and Ling, 1996). The Calabar bean, Physostigma venenosum, was once used by West Africans in a “trial by ordeal” (O’Brien, 1967). Study of the action of its active anticholinergic ingredient, the CB physostigmine (eserine), helped to establish the roles of ACh and AChE in the nervous system (Engelhart and

Table 68.1  Relative RBC AChE Levels of Adults of Selected Speciesa Species

Sex

AChE level (%)

Human

Male/female

100  8.7

Cow

Female

87.6  1.9

Guinea pig

Male/female

32.7  3.5

Horse

Male/female

28.8  9.0

Rabbit

Male/female

21.7  5.3

Rat

Male

12.6  3.0

Cat

Male/female

30b

a Manometric assay at N  3. Mean human AChE was 2180 l CO2/ 30 min/mg nitrogen. ACh substrate. b N  2. Adapted from Zajicek (1957).

Loewi, 1930). Other examples of naturally occurring anti-ChE agents are fasciculin from elapsid snake venom (Marchot et al., 1998), chaconine and solancine from tubers and nightshades (Nigg et al., 1996), and huperzine from moss (Patocka, 1998). The association of cholinergic transmission with Alzheimer’s disease served as a stimu­ lus for modern studies of natural anti-ChE agents (Francis et al., 1999; Nordberg and Svensson, 1998), part of the search for treatments of this common disorder. Tissue ChEs may have specific but still unknown roles in addition to their regulation of neural transmis­ sion. For example, there is evidence (Anderson and Key, 1999; Chiappa and Brimijoin, 1998; Layer et al., 1998; Robitzki et al., 1997; Sharma and Bigbee, 1998; Sperling et al., 2008) for a developmental function for ChEs based on studies of neurite outgrowth in retinal and dorsal root ganglion cultures and embryos using immunological, sense and antisense oligonucleotides and inhibitors. The functions of AChE and BuChE may overlap. The suc­ cessful development and survival of a knockout mouse mutant lacking AChE indicates that other enzymes such as BuChEs can function in its stead (Li et al., 2000; Xie et al., 2000). There have been persistent reports concerning pesticideinduced ocular damage (recognized clinically as Saku dis­ ease in Japan). These studies, reviewed by Dementi (1994) and Jaga and Dharmani (2006), have stimulated stud­ ies by the U.S. Environmental Protection Agency (EPA) (Atkinson et al., 1994; Boyes et al., 1994), although with­ out striking results. However, there are reports of visual changes and damage during growth and development (Geller et al., 1998; Wyttenbach and Thompson, 1985). Hamm et al. (1998) found that diazinon, a widely used pesticide, damaged the development of the neural retina in Medaka, a fish.

Chapter | 68  Cholinesterases

1459

Table 68.2  Plasma Hydrolysis of Choline Esters of Selected Species of Mammalsa Species

ACh

PrCh

BuCh

MeCh

BzCh

Human

135

310

360

 2

60

Cow

  3

  4

  2

 0

 0

Guinea pig

  50

130

170

 5

20

Horse

130

225

365

 2

40

Rabbit

  16

  16

  10

 4

 2

Rat

  20

  30

  15

15

10

Dog

  70

115

180

 6

30

Cat

  50

  75

150

 5

12

ACh, acetylcholine; BuCh, butyrylcholine; BzCh, benzoylcholine; MeCh, methylcholine; PrCh, propionylcholine. a Manometric assay.  lCO2/0.1 ml plasma/30 min. N  3 or more animals. Source: Augustinsson (1959a).

Table 68.3  Plasma Hydrolysis of Choline Esters of Selected Species of Birds, Reptiles, Amphibians, and Fisha Species

ACh

PrCh

BuCh

MeCh

BzCh

Chicken

37

  71

36

19

 2

Duck

43

  74

67

 7

 8

Turtle

14

103

27

 7

 1

Rana

40

  90

87

 2

10

Xenopus

 2

  5

 9

—

—

Pike

12

  2

 1

 1

 1

ACh, acetylcholine; BuCh, butyrylcholine; BzCh, benzoylcholine; MeCh, methylcholine; PrCh, propionylcholine. a Manometric assay.  lCO2/0.1 ml plasma/30 min. N  3 or more animals. Source: Augustinsson (1959b).

68.3  Substrate preferences and selective inhibitors AChEs prefer ACh as a substrate. Substrate preferences and activities of BuChEs vary with the species. For exam­ ple, rat plasma BuChE activity has been reported to favor propionyl rather than butyryl substrates, and cows have been shown to have hardly any plasma ChE activity at all (Tables 68.2 and 68.3). An important distinction between the AChEs and BuChEs is their response to substrate concentration. AChEs are inhibited by substrate in excess of a few millimoles; BuChEs are less sensitive (Hoffmann et al., 1989; Wilson, 1999). In general, mammalian and avian AChEs rapidly hydrolyze ACh and its thiocholine analog acetylthiocholine (ATCh). Mammalian, but not avian, AChEs preferentially hydrolyze acetyl--methylcholine. AChEs are selectively inhibited by several agents, including the CB BW284c51 (1,5-bis(4-allyl-dimethylammoniumphenyl)pentan-3-one

dibromide) (Austin and Berry, 1953; Holmstedt, 1957; Silver, 1974) and ARA 1327 (Augustinsson et al., 1978). BuChEs are preferentially inhibited by iso-OMPA (tetraisopropylpyrophosphoramide; Austin and Berry, 1953), ethopropazine (Mikalsen et al., 1986), and quinidine (Wright and Sabine, 1948). Effective concentrations for these selective inhibitors may vary by species. Useful start­ ing points for testing are 0.1–0.01 mM.

68.4  Multiple molecular forms   and life history ChEs are polymorphic proteins, occurring in multiple mole­ cular asymmetric and globular forms (Figures 68.1 and 68.2). The asymmetric forms tend to be localized at synapses and motor endplates. They have glycosylated heads joined by sulfhydryl groups and collagen tails. The heads contain the active sites; the collagen tails attach the enzymes to cell

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1460

A I2

s s ss

Hs

H

s s

G1 s s

s s s s

s s

A8

s ss s

H

s s s s

s s

s s

s s

G2

s s

s s ss

s s

H

s s

s s s s ss s s

s s

H

G4

H H

A4

Figure 68.1  Organization of subunits in the molecular forms of ChEs. Each circle represents a catalytic subunit. Globular forms are represented by “G.” Asymmetric forms are designated by “A.” Linear elements rep­ resent collagen-like tails. Disulfide bridge locations from studies of eel electroplax AChE, assumed to reflect ChEs from many sources (after Brimijoin, 1992).

Nucleus

surfaces. The globular forms make up the catalytic subunits (Taylor, 1996). These forms have similar kinetic properties but differ in their ionic and hydrophobic interactions. ChE forms (Massoulie et al., 1993, 1999) are synthe­ sized as catalytic globular monomers (G1) that oligomer­ ize via disulfide bonds into multiple G2 and G4 forms (see Figure 68.1). The G1 subunits are synthesized within cells (e.g., nerve, muscle, and liver), glycosylated, and then secreted. Collagen tails are attached to one, two, or three catalytic tetramers to yield A4, A8, and A12 asymmetric forms. Collagen-tailed forms become attached to the cell surface at specific binding sites. Globular forms are released into body fluids or bind to cell surfaces through hydrophobic amino acid sequences or glycophospholipids (Taylor, 1996). Antibodies have been prepared to several purified AChEs and BuChEs; specific protein and nucleic acid sequences have been determined and altered by site-directed mutagenesis (Doctor et al., 1998). AChE and BuChE forms are each coded by single genes. The gene coding sequence for the collagen-tailed forms con­ tains a C-terminal extension of 40 amino acids, the T-peptide, that interacts by a short proline-rich attachment domain with the collagen. A single collagen gene (ColQ) is associated

Synthesis, sequestering glycosylation, assembly of globular forms, degradation

AChE mRNA RER 20–30%

Degradation 70–80%

Processing, assembly of asymmetric forms

Golgi Transport Vesicles Plasma membrane

Membrane bound

Extracellular matrix

Secretion, membrane incorporation and turnover

Binding in basal lamina

Extracellular fluids Figure 68.2  Life history of ChEs. The enzyme is synthesized as a monomeric globular form (G1). Up to 80% of the enzyme is degraded by intra­ cellular proteases after ribosomal translation and before transit of the Golgi. In the Golgi, secretory forms (black triangles) are sequestered from membrane-bound forms (open triangles), collagen-like tails are added to asymmetric forms, the peptide backbone is glycosylated, and the ChE becomes enzymatically active. Globular secretory forms may escape the synaptic cleft to enter extracellular fluids, blood, or external secretions. Asymmetric forms are probably bound quantitatively by adsorption or entrapment in the polyanion-rich, fibrous matrix of the extracellular synaptic basal lamina (after Brimijoin, 1992).

Chapter | 68  Cholinesterases

1461

with both the AChE and the BuChE collagen-tailed forms. The globular forms are synthesized by an alternative splice variant, the H-exon, that is expressed instead of the T-exon [see Krejci (1998) for a model and Massoulie et al. (1999) for further information]. The three-dimensional structures (Figure 68.3) of ChEs are subjects of intense investigation (Doctor et al., 1998; Massoulie et al., 1999). In the past, the active site of AChE was loosely described by models based on the work of Nachmansohn and Wilson (1951), in which there was a negatively charged “anionic” site and an “esteratic” site of catalytic residues. The positively charged choline moiety of ACh was hypothesized to bind to the negatively charged anionic site. A nucleophilic group, assumed to be a serine residue at the esteratic (acylation) site, was proposed to catalyze the hydrolysis [see Silman and Sussman (1998) for a more detailed historical perspective]. The crystallization of ChE proteins such as the dimeric AChE form of Torpedo californica has led to a more detailed understanding of the form/function relationships of the enzyme (Doctor et al., 1998). An important feature is the embedding of the active site of AChE in a “gorge” lined with 14 aromatic residues, approximately 20 Å from the surface of the protein (see Figure 68.3). The quaternary nitrogen of cho­ line binds through interactions with  electrons of tryptophan residues (Sussman et al., 1991) at the “peripheral site” at the mouth of the gorge, a region conceptually corresponding to the historical concept of an “anionic site.” One current view of the molecular mechanism has the ester substrate led down

the gorge by molecular interactions to become hydrolyzed at the bottom by a catalytic triad of glutamate 334, histidine 447, and serine 203 residues (Doctor et al., 1998; Taylor, 1996). How the products escape from the gorge is a matter of specu­ lation; one hypothesis is that they are ejected via a “side” or “back” door, making room for the next substrate molecule; another is that they are rapidly moved to the entrance of the gorge. However, Silman and Sussman (2008), in an excel­ lent review of the subject, note that “the current view of most kineticists working on AChE would be that they could model the kinetic behavior of the enzyme without requiring a back door” and that “experimental evidence is thin” (p. 7). The amino acid sequences around the entrance to the gorge that comprise the “peripheral site” may be important in deter­ mining the differences in substrate specificity and inhibition by excess substrate between AChE and BuChE forms (Doctor et al., 1998; Reiner et al., 1999). Figure 68.3 provides a threedimensional diagram of an AChE molecule. [Other depictions may be found in Silmann and Sussman (2008) and on the Internet by searching for “acetylcholinesterase.”]

68.5  Mechanism of hydrolysis 1

124

72 124

74

74

286

86

86 282

282 297

297

283

203

341 447 295



where E is the enzyme, AX is the substrate (ACh) or inhibitor, EAX is the reversible enzyme complex, X is Ch

72 286

K

K

3 2 E  AX ←kk1 → EAX   → EA  X   →E  A

341 449

447

449

295 334

334

Figure 68.3  Three-dimensional view of the active center of AChE. Modeled from Torpedo AChE with the addition of amino acid side chains of the mammalian enzyme. Amide backbone shown by the ribbons. The catalytic triad of the enzyme is Glu334, His447, Ser203, with hydrogen bonds indicated by dotted lines. The acyl pocket is Phe295 and Phe297; the choline subsite is Trp86, Glu202, and Tyr337; the peripheral site is Trp286, Tyr72, Tyr124, and Asp74. Tyrosines 341 and 449 may help to stabilize some ligands. The catalytic triad, choline subsite, and acyl pocket are at the base of the 18–20 Å gorge; the peripheral site is at its lip. (One way to see the diagram in three dimensions is to use 3-D glasses; another is to move the page toward your eyes until the images superimpose, keeping your head level. If there appear to be three images, concentrate on the middle one.) (After Taylor, 1996.)

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(choline), and ks are reaction rate constants (Rosenberry et al., 1998; Taylor, 1996). The kinetics of ACh hydrolysis is a complicated multi­ step process (Figure 68.4 is one depiction) discussed here in abbreviated form. The first step is a nucleophilic attack of the carbonyl carbon resulting in the formation of a reversible enzyme–substrate complex (EAX), acylation of the catalytic site (EA), and liberation of choline. This is followed by a rapid hydrolysis of the acylated enzyme, pro­ ducing acetic acid and regenerating the enzyme (E  A). A similar reaction scheme applies to BuChEs (Taylor, 1996). The real-time kinetics of the enzyme reactions has been described by one biochemist (Quinn, personal com­ munication) as “approaching catalytic perfection.” The rate of ACh turnover (kcat/Km) is extremely rapid, where kcat  k2k3/(k2  k3) (the geometric mean of the two rate constants). AChE is capable of hydrolyzing 6  105 mol­ ecules of ACh per molecule of enzyme per minute, a turn­ over time of 150 s (Taylor, 1996). The upshot of this rapid rate of hydrolysis is that k1 becomes the rate-limiting step for hydrolysis of ACh and its analog ATCh. Perhaps this

Glu 334

Glu 334

Ser 203 His 447

represents the movement of substrate to the active center. The deacylation step (k3) is considered rate limiting to car­ bamoylating and phosphorylating agents. A major distinction between AChEs and BuChEs is the inhibition of AChE activity with increasing substrate concen­ tration [S]. A plot of activity versus [S] for ACh, ATCh, and, in mammals, acetyl--methylcholine, yields curves with a maximum at 1–3 mM, whereas BuChE activity increases with [S] to at least 10 mM. These effects are illustrated in Figure 68.5 from data of Wilson et al. (1997) for the human. The late Dr. A. R. Main (Hoffmann et al., 1989) noted, “Because of this inhibition, methods for determining AChE activities should employ substrate concentrations at or below [S]opt.” Unfortunately, his advice has not always been followed. The phenomenon of inhibition with excess sub­ strate may be due to interactions at the peripheral site. Rosenberry et al. (1998) present evidence that excess sub­ strate inhibition and the action of peripheral site inhibitors such as propidium are brought about by the imposition of a steric blockade in the catalytic pathway. Figuratively, a chemical cork blocks the gorge.

Glu 334

Ser 203 His 447

Glu 334

Ser 203 His 447

Ser 203 His 447

Gly 122 Gly 121

A Enzyme-substrate complex Glu 334

B Tetrahedral intermediate

Trp 86

Glu 334

Ser 203 His 447

Trp 86

Trp 86

C Acetyl enzyme Glu 334

Ser 203 His 447

Glu 334

Ser 203

Trp 86

Trp 86

F Neostigmine

Glu 334

Glu 334

Ser 203

Glu 334

His 447

His 447

Trp 86

H Dimethyl carbamoyl enzyme Glu 334

Ser 203

His 447 Gly 122 Gly 121

Ser 203

Trp 86

G Hydrolysis of dimethyl carbamoyl enzyme

Ser 203

Trp 86

His 447

His 447

E Edrophonium

D Hydrolysis of acyl enzyme

Ser 203

His 447 Gly 122 Gly 121

Gly 122 Gly 121 2PAM

l Diisopropylfluorophosphate

MOLECULE

Trp 86

J Diisopropyl phosphoryl enzyme

carbon

oxygen

Trp 86

K Aged phosphoryl enzyme

nitrogen

hydrogen

Trp 86

L Reactivation of phosphoryl enzyme

phosphorus

Trp 86

fluorine

Figure 68.4  Hydrolysis of ACh by AChE and inhibition and reactivation of the enzyme. (A) Binding of ACh. (B) Attack by the serine hydroxyl and formation of a transient tetrahedral intermediate. (C) Loss of choline and formation of the acylated enzyme. (D) Deacylation of the enzyme by H2O attack. (E) Binding of the reversible inhibitor edrophonium. (F) Binding of neostigmine. (G) Formation of the carbamoylated enzyme. (H) Hydrolysis of the carbamoylated enzyme. (I) Binding of diisopropyl fluorophosphate. (J) Formation of the phosphoryl enzyme. (K) Formation of the aged form of the phosphoryl enzyme. (L) Attack by pralidoxime (2-PAM) to regenerate active enzyme (after Taylor, 1996).

Chapter | 68  Cholinesterases

1463

The knowledge of the three-dimensional structure of the ChEs has led to a better understanding of the mecha­ nisms of action of drugs and chemical agents that inhibit the hydrolysis of choline esters. In general, there are three major domains for inhibitors to bind: the acyl and choline pockets of the active center and the peripheral anionic site. Taylor (1996) uses edrophonium and tacrine as examples of reversible inhibitors that bind to the choline subsite near tryptophan 86 and glutamate 202; other reversible inhibitors, such as fasciculin and propidium, bind to the peripheral anionic site on AChE at the lip of the gorge encompassed by tryptophan 286 and tyrosines 72 and 124. CBs and OP pesticides inhibit enzyme activity by acting as alternate substrates to ACh. CBs give rise to a carbamoylated enzyme that is more stable than the acylated enzyme, taking minutes instead of milliseconds to rehydrolyze. OPs are true hemisubstrates; they covalently bind with the ser­ ine at the active center, forming a tetrahedral configuration that resembles the transition state formed during hydrolysis of ACh. If the alkyl groups on the OP are methyl or ethyl, spontaneous regeneration may require hours or even lon­ ger if tertiary alkyl groups are involved. Loss of one of the alkyl groups, a phenomenon known as aging, further stabi­ lizes the phosphorylated enzyme, permanently inhibiting its catalytic ability. Thus, it is not appropriate to use the terms “reversible” and “irreversible” to refer to the inhibitions brought about by CBs and OPs, respectively. Both classes of chemicals react covalently with the active center of the enzyme, and at some stage of the reaction sequences both enzyme-inhibitor complexes are rehydrolyzable.

The toxicities of OPs and CBs are often correlated with the extent of their inhibitions of brain AChE. Figure 68.6 depicts the relationship between the toxicity in vivo of 30 directly acting OPs and their inhibition of AChE in vitro. However, such relationships do not necessarily mean there are simple relationships between inhibition of AChE activity in an organ or tissue and in the test tube. For example, Mortinsen et al. (1998) measured IC50 values for chlorpyrifos-oxon with tissues from 4-day-old and adult rats (Table 68.4). The IC50s from young and adult brains were similar; the IC50s from the other tissues were not, dif­ fering by 5:5- (liver) and 20- (plasma) fold even though the IC50s of immunoprecipitated purified AChEs were the same, regardless of tissue or age. Possible interference of BuChE activities was excluded by using the specific

1,000 500

1

200 2

100 50

LD50 (mg/kg)

68.6  Toxicities of anticholinesterases

20

3

31

4

10

5

6, 7, 8

5

10

9 11

2 32

1

12 13

14

15

17,18

0.5

19

20 22

6

21 23

0.2 25

ChE activity (µmol/min/ml)

5

0.1

1 2

4

6

16

8

24 26

10

12

Pl50 4

3

2 pH 8.0 pH 7.2

1

0 0.1

1

10

[Acetylthiocholine] (mM) Figure 68.5  Effect of substrate concentration on activity of human blood AChE at pH 8.0 and 7.2. Ellman assay, pooled blood (after Wilson et al., 1997).

Figure 68.6  Toxicity in vivo of directly acting OPs versus their inhibi­ tion of AChE in vitro. (1) Dipterex. (2) O,O-diethyl-4-chlorophenylphos­ phate. (3) O,O-diethyl-bis-dimethyl pyrophosphorodiamide (sym). (4) TIPP. (5) O,O-diethylphosphostigmine. (6) Isodemeton sulfoxide. (7) Isodemeton. (8) Isodemeton sulfone. (9) DFP. (10) Diethylamidoethoxyphosphoryl cya­ nide. (11) O,O-dimethyl-O,O-diisopropyl pyrophosphate (asy). (12) Diethyl; amido-methoxy-phosphoryl cyanide. (13) Tetramethyl pyrophosphate. (14) O,O-diethyl phosphorocyanidate. (15) O,O-dimethyl-O,O-diethyl pyrophos­ phate (asym). (16) Soman. (17) TEPP. (18) O-isopropyl-ethylphosphone-flu­ oridate. (19) Tabun. (20) Amiton. (21) Diethylamido-isopropoxy-phosphoryl cyanide. (22) O,O-diethyl-S-(2-diethylaminoethyl)phosphorothioate. (23) Sarin. (24) O,O-diethyl-S-(2-triethylammoniumethyl)thiophosphate iodide. (25) Echothiophate. (26) Methylfluorophosphorylcholine iodide. (27) Methylfluorophosphoryl--methylcholine iodide. (28) O-ethyl-methylphos­ phorylthiocholine iodide. (29) Methylfluorophosphoryl-homo-choline iodide. (27), (28), and (29) with LD50 values of 0.03–0.07 mg/kg are not shown. (31) Schradan and (32) dimefox (shown in the graph) were not used to calculate the regression (after Holmstedt, 1963 as shown in Gallo and Lawryk, 1991).

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1464

Table 68.4  Tissue IC50 Values for 4-Day-Old and Adult Mice after Chlorpyrifos-Oxon

Table 68.5  Necrosis of Rat Muscle after DFP and Botulinum Toxina

Tissue

Neonate (nM; mean  SE)

Adult (nM; mean  SE)

Treatment

EDL

SOL

Saline

0

0

Brain

9.6  0.1

10  0.2

DFP

85.5

28

Liver

96  3

530  50

Btx

0

0

Plasma

18  1.5

330  28

DFP  Btx

0

1.59

Purified

3 nM for all tissues

Source: Mortinsen et al. (1998).

BuChE inhibitor iso-OMPA. A-esterase destruction, car­ boxyesterase binding, and sequestration of the lipophilic OP were considered possible factors accounting for the dif­ ferences between the in vitro and in vivo findings. One example of a carboxyesterase is serum paraox­ onase (PON1), an A-esterase associated with high-­density plasma lipoproteins; PON1 destroys OPs such as the oxon analogs of parathion and chlorpyrifos. Direct evidence for its role in detoxifying OPs was provided by the demon­ stration that mice exposed to chlorpyrifos were protected against cholinesterase inhibition and toxicity by admin­ istration of purified PON1 (Li et al., 1995). Shih et al. (1998) demonstrated that knockout PON1-deficient mice were more sensitive to chlorpyrifos and chlorpyrifos-oxon than were genetically unaltered mice. Blood ChEs also have been shown to protect animals from OP toxicity. Studies on chemical warfare agents, led by the initial report of Wolfe et al. (1987), have shown that injection with puri­ fied AChE can protect mice and other animals from expo­ sure to OPs (Doctor et al., 1991). Many physiological actions of anti-ChEs are those expected from an excess of ACh caused by the inhibition of its catalysis. Specific symptoms depend on the chemi­ cals and the receptors concerned (discussed elsewhere in this handbook). Early signs of cholinergic poisoning likely involve stimulation of muscarinic neuroeffectors of the parasympathetic system. Symptoms include slowing of the heart (bradycardia), constriction of the pupil of the eye (miosis), diarrhea, urination, lacrimation, and saliva­ tion (Spencer et al., 2000; Taylor, 1996). Overstimulation at skeletal nicotinic neuromuscular junctions (motor end­ plates) causes muscle fasciculation (disorganized twitch­ ing) and, at higher doses, muscle paralysis. Increased ACh at cholinergic junctions of the sympathetic and ­parasympathetic autonomic ganglia affects the eye, blad­ der, heart, and salivary glands. Finally, anti-ChEs affect junctions of the CNS, producing hypothermia, tremors, headache, anxiety, convulsions, coma, and death. Whether or not there are consistent behavioral effects at low dose levels of OPs and CBs, such as deficits in learning and memory, is a matter of ongoing research.

Btx, botulinum toxin; DFP, diisopropyl fluorophosphate; EDL, extensor digitorum longus; SOL, soleus muscle. a Necroses/1000 fibers. DFP injected 48 h after Btx; sampled 24 h later – 1.5 mg/kg sc. Source: Sket et al. (1991).

In addition to affecting the nervous system, the excess ACh brought about by anticholinergic agents can cause a transient myopathy (Dettbarn, 1984). In vivo studies of Meshul (1989) and in vitro studies of the late Miriam Salpeter and colleagues using ACh receptor antagonists (Leonard and Salpeter, 1979) show that this is due to an influx of Ca2 and other cations into the postsynaptic cell. For exam­ ple, necrosis due to diisopropyl fluorophosphate (DFP) was prevented in vivo by treatment with -bungarotoxin, a snake venom agent that binds irreversibly to the nicotonic ACh receptor (Kasprzak and Salpeter, 1985). Further evi­ dence that the necrosis is ACh mediated was provided by Sket et al. (1991), who demonstrated that botulinum toxin (a presynaptic inhibitor of ACh release) prevented muscle necrosis induced by DFP in the rat (Table 68.5). ACh-induced myopathy may cause necrosis in 10–30% of the muscle fibers around the motor endplates (Dettbarn, 1984). Prolonged muscle weakness and muscle damage lasting several weeks or longer may occur. A similar tran­ sient muscle damage in humans has been termed interme­ diate syndrome (Senanayake and Karalliedde, 1987). Although it is generally accepted that most of the effects of OPs and CBs are due to inhibition of AChE, there is evi­ dence for other modes of action of these agents. Anti-ChE pesticides have been shown to directly affect pre- and post­ synaptic events (Pope, 1999). Electrophysiological studies suggest that choline may act as a regulator of nicotinic recep­ tors in the CNS (Albuquerque et al., 1998; Alkondon et al., 1997). Malathion and other OPs have been shown to affect the immune responses of mammals and fish (Beaman et al., 1999; Rodgers and Ellefson, 1990; Rodgers and Xiong, 1997). A few OPs, including some pesticides (e.g., isofenphos and chlorpyri­ fos) and at least one chemical warfare agent (sarin), have been shown to cause OPIDN, a distal axonopathy. Inhibition of ChEs plays a role in drug interactions. For example, cocaine is both detoxified by and is itself a reversible inhibitor of BuChEs. Studies on experimental animals indicate that depressing ChEs with anti-ChE treatments intensifies the toxic effect of cocaine (Hoffman et al., 1992).

Chapter | 68  Cholinesterases

Genetic variation between individuals can play an impor­ tant role in the toxicity of anti-ChEs. One example is humans with inherited low levels of plasma BuChEs. Although usu­ ally symptomless, patients with genetically low BuChE given succinylcholine (or a similar drug) during surgery to induce relaxation of muscles are unable to speedily destroy the drug, intensifying and prolonging its activity, sometimes with seri­ ous consequences. People with such a genetic makeup can be detected by assays using dibucaine and fluoride (Silk et al., 1979). At least two genetic polymorphisms, those of low BuChE and PON1, are considered risk factors for OP and CB pesticide exposures (Shih et al., 1998).

68.7  Assay techniques The history of ChE assays has been reviewed (Hoffmann et al., 1989; Silver, 1974; Wills, 1972; Wilson, 1999; Witter, 1963). Some techniques are “endpoint” assays; others record the time course of hydrolyses. The colorimetric and radiomet­ ric techniques currently used in my laboratory are based on those of Ellman et al. (1961) and Johnson and Russell (1975) and found in Current Protocols in Toxicology (Wilson and Henderson, 2007). One early assay to determine the hydro­ lysis of acetylcholine used a Warburg manometer to mea­ sure the CO2 released from a bicarbonate containing buffer (Ammon, 1933). The particulars of this method are outlined by Wills (1972). Although accurate, it is little used today. Following World War II, there was widespread develop­ ment of OPs for pest control. A veritable OP race ensued. By 1950, scientists were seeking rapid, accurate, and convenient clinical assays for blood ChE levels. Metcalf (1951) expressed the rationale for monitoring: “Since the cholinesterases of human blood are very sensitive to the presence of cholinesterase inhibitors, it appears that peri­ odic estimation of blood cholinesterase levels may pro­ vide an indication of dangerous levels of overexposure to these toxicants.” Several of the methods developed a half century ago are still in use today. Hestrin (1949) determined the ACh remaining after incubation by reacting it with hydroxyl­amine under alkaline conditions to form a reddishpurple complex read at 515 nm. Metcalf (1951) adapted the method for drops of blood obtained with a spring-loaded lancet. Okabe et al. (1977) oxidized the choline released during ACh hydrolysis; the hydrogen peroxide produced was determined with an indicator reaction at 500 nm (Abernathy et al., 1988). Three kinds of ChE assays are commonly used today utilizing electrometric (pH), radiometric, and colorimetric techniques. Specific examples are listed in Table 68.6.

68.7.1  Radiometric An example of a radiometric technique is that of Johnson and Russell (1975). It is based on the differential solubility of ACh and its hydrolysis products in organic and ­aqueous

1465

Table 68.6  Common ChE Assays Test

Basis

Conditions

Analysis

Johnson and Russell (1975)

Radiometric [3H]ACh

Endpoint

Micro, fast, costly disposal

Michel (1949)

pH

Rate, ACh

Simple, slow, cheap

Ellman et al. (1961)

Colorimetric

Rate, ATCh

Micro, rapid

Adapted from Wilson and Henderson (1992).

media. Sample and tritiated ACh are reacted together; an organic solvent is added when the assay is completed. The unhydrolyzed radioactive ACh substrate remains in the aque­ ous phase, quenching its scintillation. The hydrolyzed radio­ active acetate moves into the organic phase and is counted. The degree of substrate hydrolysis is determined by compar­ ison to a nonhydrolyzed ACh blank and a totally hydrolyzed ACh sample (usually accomplished by incubation with an excess of electric eel AChE). This and similar assays have high sensitivity, problems with dilution of samples encoun­ tered with “reversible” carbamate chemicals are minimized, many tubes may be measured at once, the sample size is small, and readouts may be computerized. However, radio­ active assays involve a high initial investment and costly dis­ posal of radioactive waste. Being an endpoint assay, many duplicate samples are needed to run a kinetic analysis. Potter et al. (1993) applied a radioactive assay similar to that of Thomsen et al. (1988, 1989) in a field study of applicators of OPs and fumigants. In this radiometric endpoint method, 14 C-labeled ACh hydrolysis was stopped by ethanol/glacial acetic acid, the labeled acetic acid was evaporated, and the unhydrolyzed ACh was counted.

68.7.2  pH The modified Michel method (Michel, 1949) directly deter­ mines the change in pH due to ACh hydrolysis with a pH meter or by titrating the acetic acid produced with NaOH while keeping the pH constant (Groff et al., 1976; Nabb and Whitfield, 1967). Potentiometric methods are reliable, they use simple reagents, and they are relatively inexpen­ sive. However, they are limited by their relative insensitiv­ ity; they often have larger sample requirements and lower outputs than radiometric methods. Several micro-pH meth­ ods have been described (Gage, 1967); one using 10 l of capillary blood was described by Mosca et al. (1995). pH assays can have relatively low variability. Results from an early pH assay study by Rider et al. (1957) of 12 males and 12 females aged 40 years, taken from a study of 800 donors at a San Francisco blood bank, are shown in Table 68.7. Whether or not there is a difference between male and female subjects (as suggested in Table 68.7) is

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Table 68.7  Plasma and RBC ACh Hydrolysis Levels for 40-Year-Old Blood Donorsa Men

Women

RBC ChE

0.766  0.081

0.750  0.082

Plasma ChE

0.953  0.157

0.817  0.187

a

pH/h/0.02 ml RBCs or plasma, 25°C, mean  SD. Source: Rider et al. (1957).

not clear. Many studies have noted a higher variability of plasma than RBC ChE activities. Gage (1967) points out that “population averages obtained by different investiga­ tors … depend to some extent upon the method of assay used.” Even so, Gage concluded on the basis of the studies available to him that “an individual in good health with a plasma ChE 33% below the population average, or a red cell cholinesterase more than 20% below, has an abnor­ mally low value and has probably been exposed to a ChE inhibitor.” Such “red-alert level” estimates have changed little in more than 30 years.

68.7.3  Thiol Substrates and the Ellman Assay Thiocholine substrate assays based on the work of Ellman et al. (1961) may be the most popular of ChE assays, replac­ ing the pH methods. Many variations of the original Ellman assay have been published. The conditions of three popular commercial assays are shown in Table 68.8. Several auto­ mated versions (e.g., those of Technicon and COBIAS) are no longer marketed. The basis for the thiocholine assays is the hydrolysis of ATCh or related substrates by the enzyme, producing a thiol group that reacts with a sulfhydryl-sensitive chromogen such as dithiobisnitrobenzoate (DTNB). In this case, peak absorption of the thionitrobenzoate produced is at 410–412 nm. The original Ellman assay used cuvettes; some laboratories have modified the assay for a 96-well microplate reader in a manner similar to that described by

Doctor et al. (1987). The Boehringer–Mannheim (Roche) Diagnostics kits (1981), Nos. 124117 and 450035, use ATCh as a substrate; the Sigma Diagnostic Cholinesterase (PTC) kit Procedure No. 422 uses propionylthiocholine (Sigma Diagnostics, 1989) for both AChE and BuChE. The Boehringer–Mannheim version used with a Hitachi automated spectrophotometer reads the reaction at 480 nm. This avoids possible interference of the Soret band of hemoglobin (Hb), but at the expense of reducing the sen­ sitivity of the assay (Wilson and Henderson, 1992). The manual instrument kits recommend 405 nm. Instructions for both Boehringer–Mannheim and Sigma kits are for human blood. The reader is cautioned that the kits may not be suited to the needs of clinical veterinary laboratories or researchers without modification. In the case of the human, Wilson et al. (1997) showed that the high substrate concen­ tration and low pH of the Boehringer–Mannheim kit intro­ duce a difference of 40% in the manual assay compared to the Ellman assay run under optimum conditions (see Figure 68.5). This is because the optimum concentration of sub­ strate for human AChE is 1 or 2 mM and the optimum pH for the assay is 8.0, whereas the kit uses a substrate con­ centration of 5.4 mM and a pH of 7.2. Absorbances would have been even more reduced if the measurements for the Boehringer–Mannheim substrate and pH conditions were read at 480 nm (Wilson and Henderson, 1992). There have been recommendations that the Ellman assay be modified to use dithionicotinic acid, a chromo­ gen that absorbs in the near-ultraviolet at 340 nm, to avoid the interference of Hb at 410 nm, thus divorcing the wave­ length of the assay from the Soret band of Hb (Christenson et al., 1994; Loof, 1992; Willig et al., 1996). However, instruments reading in the near-ultraviolet are often more costly than those that register in the visual range. The sta­ bility of modern instruments should be sufficient to over­ come the increased noise level of the assay at 410 nm, providing activity levels are sufficiently high. The sub­ strate and pH of the Sigma Diagnostics Kit are optimal for neither RBC nor plasma BuChE. Augustinsson et al. (1978) proposed using propionylthiocholine, the Sigma

Table 68.8  Common Thiocholine-Based Assays Boehringer–Mannheim (Roche) Parameter

Ellman

Manual

Auto

Sigma

Wavelength (nm)

412

405

480

405

Substrate (mM)

ATCh, 0.5–1.0

ATCh, 5.4

ATCh, 5.4

PrTh, 5.0

pH

8.0

7.2

7.2

7.2

DTNB (mM)

0.32

0.24

0.24

0.25

PrTh, propionylthiocholine. Source: Wilson (1999).

Chapter | 68  Cholinesterases

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kit substrate, as a compromise substrate for both enzymes. Under the conditions of their assay, it was no better a sub­ strate for one than it was for the other. They recommended measuring the reaction in the ultraviolet, avoiding Hb interference that might well play a role, given the reduced sensitivity of the assay. The Sigma kit is no doubt excel­ lent as a screen for patients with reduced BuChE activities before they undergo surgery and treatment with muscle relaxants such as succinylcholine and mivacurium, but its use may be more difficult to justify, given the excellence of today’s instrumentation, for the determination of RBC AChE activities to detect exposures to pesticides because the conditions are not optimum for the enzyme. What to do? With blood enzyme assays, no one size seems to fit all. One approach is to focus on ATCh hydrolysis, and spe­ cific inhibitors such as iso-OMPA or quinidine to inhibit BuChE, establishing an estimate of BuChE by difference, accepting, for convenience, the use of an inappropriate substrate and substrate concentration. The problems of using commercial kits with condi­ tions designed for the human may be exacerbated when the kits are applied to other species. Many studies (some of which are summarized in Tables 68.9 and 68.10) indi­ cate that, with other species, the relative activities of AChE and BuChEs, and even the properties of BuChEs, may dif­ fer from those of human enzymes. Lack of consideration for this may result in possible discrepancies and misinfor­ mation. For example, Harlin and Ross (1990) used adult bovine blood to establish conditions for the determination of cholinesterase activities in the only approved AOAC assay with ATCh as substrate and bovine blood. This excellent round-robin study did not discuss the report of Augustinsson (1959a, Table 51.2) that bovines had hardly any plasma cholinesterase activity. Further study is needed to establish whether the assay was determining only RBC AChE and is useful for assaying it when serum cholinester­ ase is low or absent. Another problem arises when studies are undertaken with species that lack RBC AChE; the liter­ ature examined by the author suggests that only mammals have RBC AChE and that the properties of serum ChEs may vary with the species (see Table 68.9). A further problem arises when, as is common in studies of laboratory animals that play a role in pesticide registration,

the investigators assume AChE activity is restricted to the RBCs. Table 68.10 illustrates the wide differences between AChE activity in blood for humans and for two species often used in registration-geared research – the dog and the rat. Of the three, only the human has a relatively low ATCh hydrolysis rate in the plasma. Indeed, the AChE activity of rat plasma (Table 68.11) may exceed the activity of plasma BuChE. Another source of plasma AChE is the platelets (Table 68.12). Although the AChE activity per platelet is high, its relative contribution to blood ChE is low because the plate­ let content of blood is several orders of magnitude less than its content of RBCs. Studies of the particular AChE forms in the plasma of many species are lacking. An important and often unrecognized problem with thiocholine-based assays is the presence of a transient nonlinear “thiol oxidase” reaction with the color reagent

Table 68.9  Relative Substrate Specificity of Plasma ChEsa Species

Propionyl

Butyryl

Benzoyl

Dog

150

253

60

Horse

161

231

28

Cat

111

211

27

Human

155

192

36

Duck

139

153

25

Squirrel

122

144

14

Ferret

122

139

28

Hamster

153

128

24

Rat

211

119

17

Chicken

147

83

6

Mouse

139

75

11

Butyryl favoring

Propionyl favoring

a

Normalized to plasma ChE hydrolysis of ACh  100%. Adapted from Hoffmann et al. (1989, Table 11.4) and Myers (1953).

Table 68.10  Relative Acetyl Ester Hydrolysis Levels of Several Speciesa Species

RBC (mean  SD)

No. of trials

Plasma (mean  SD)

No. of trials

Human

135  29

60

37  9.3

56

Dog

17.9  3.5

18

25.4  5.5

18

Male rat

9.0  1.3

24

4.3  1.0

45

a

ATCh 0.8 mM (RBC); 7 mM (plasma); Autotechnicon analyzer. Adapted from Humiston and Wright (1967, Table 4).

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Table 68.11  AChE/BuChE Activity of Rat Plasmaa Total ChE (mU/ml, mean  SE)

BuChE (mU/ml, mean  SE)

AchE (mU/ml, mean  SE)

452  17

175  12

210  8

a

Eighteen rats, Ellman assay. Total with acetylthiocholine; BuChE with butyrylthiocholine; AChE with acetylthiocholine  0.1 mM iso-OMPA. Source: Traina and Serpietri (1984).

Table 68.12  Comparison of RBC and Platelet Activities of Several Speciesa Species

Sex

Red blood cell Platelet

Human

Male/female

2180  189

0

Cow

Female

1910  42

53  21

Guinea pig

Male/female

713  76

493  119

Horse

Male/female

627  200

1500  252

Rabbit

Male/female

473  116

3930  208

Rat

Male

274  64

4240  1070

30

5450

Cat

b

Male/female

a

Manometric, lCO2/30 min/mgN; N  3. N  2. Adapted from Zajicek (1957). b

DTNB in RBCs of some species. It is necessary that assays with species such as the rat that have such high “tissue blanks” be designed to either circumvent or correct for them because they can lead to indeterminate errors greater than 70% (Table 68.13). One way is to include the appro­ priate blank in the assay. Another is to preincubate the samples for a few minutes before adding substrate until the nonlinear first 5–10 min of the assay is over (Chancy et al., 2000). Lack of consideration of such problems in clinical assays of experimental animals submitted for regu­ latory purposes may have played a role in the difficulties encountered when the U.S. EPA tried to compare data on the rat from different laboratories (U.S. EPA, 1992). Errors may be compounded. For example, a large and indetermi­ nate error will occur when the Boehringer–Mannheim kit assays are applied to rats without preincubating the sam­ ples. Such an error could be exacerbated by the low RBC activities and the misinterpreted plasma “BuChE” activities that are actually due to a mixture of the AChE and BuChE activities. The lesson is that whatever the assay, it is critical that its conditions be validated for each species, tissue, and chemical under study. CBs represent a special problem because of the ease of rehydrolysis of carbamate–ChE complexes. The sensitiv­ ity of the assay requires dilution of the enzyme samples, but the act of dilution itself promotes rehydrolysis of the

Table 68.13  RBC DTNB Background Reaction in Various Speciesa Species

% Total activity

Human

10

Monkey

18–27

Dog

30–60

Rabbit

50–70

Rat

60–75

Mouse

4

Brain (all species)

N/A

Source: Loof (1992). a Transient activity, first 5–10 min, Ellman method.

enzyme. Several investigators have suggested ways to mini­ mize the problem, including Thomsen et al. (1988) for the radiometric assays and Nostrandt et al. (1993) for Ellmantype thiocholine-based assays. Regardless of the assay used, samples should be kept iced from the time of their collection to their assay. It is unfortunate that some are under the impression that OP exposures lead to “irreversible” inhibitions, and that icing a blood sample is unnecessary as long as an anticlotting agent such as EDTA or heparin has been used. To the con­ trary, as discussed elsewhere in this handbook, ChE inhi­ bitions by methyl-OPs such as azinphosmethyl (Guthion) have relatively rapid rates of spontaneous reactivation at room temperature; in this case, the half-life of recovery of activity for azinphos-methyl is approximately 2.5 h (Wilson et al., 1992b). The lack of a requirement to keep samples on ice as a part of assay protocols makes it difficult to inter­ pret the results of otherwise excellently designed and exe­ cuted studies such as that of Yeary et al. (1993). Although the instructions that accompany the commercial ChE mon­ itoring kits discussed in this chapter recommend storing samples under refrigeration [4°C (Boehringer–Mannheim) and 2–6 or 20°C (Sigma Diagnostics)], several clinical laboratories I contacted said they did not specify that sam­ ples be delivered to them on ice.

68.7.4  Variability Regarding studies using the Ellman assay, a paraphrase from George Orwell’s classic “Animal Farm” (1945) might be that “all Ellman thiocholine assays are created equal, but some are more equal than others.” The variety of condi­ tions used in field studies and laboratory experiments with thiocholine substrates, and the lack of an accepted standard assay and enzyme, make it difficult to compare the activi­ ties obtained from one experiment to another (Carakostas

Chapter | 68  Cholinesterases

and Landis, 1991; Wilson et al., 1992a). Perhaps this is what has led to the idea that thiocholine-based ChE studies are “too variable” to be relied on for population exposure research and regulatory decisions, even though a number of carefully performed studies such as Sanz et al. (1991) indicate that ChE assays of populations can be performed with satisfactory results. For example, Sidell and Kamiskis (1975) measured RBC and plasma ChEs of a group of 22 subjects biweekly for a year using the Technicon autoana­ lyzer. They found that RBC AChE levels varied less than hematocrit, Hb, or RBC counts. The annual average range of AChE values was 8% for men and 12% for women. The corresponding plasma ChE values were 25% for men and 24% for women. In general, as is true for the pH meth­ ods discussed previously, plasma ChE values appear more variable than RBC AChE activities, whether the data are expressed on a per cell or a per Hb basis (Table 68.14). Bellino et al. (1978) used fingersticks and saponinhemolyzed human blood to carefully determine the optimum conditions for the Ellman assay for human RBCs and plasma, demonstrating inhibition of ATCh hydrolysis with excess substrate and an S-shaped substrate–concentration curve with butyrylthiocholine. They established that 0.01 mM eser­ ine inhibited AChE and 0.3 mM eserine inhibited BuChE. Similarly, they found that 7.0 mM totally inhibited both AChE and BuChE activity, and 2.8 mM of sodium dodecyl sulfate inhibited AChE but not BuChE. Their study of healthy men and women showed relatively low variability (Table 68.15). Large sample numbers were obtained in a study of farm worker families from migrant housing centers in California by Wilson et al. (1998), in which almost 900 volunteers contributed fingersticks of blood. Ten micro­liters of blood was hemolyzed at the site, transported on ice to the labo­ ratory, stored at 70°C, analyzed under optimum assay conditions for RBC AChE using quinidine to inhibit plasma BuCh, and expressed on an Hb basis. Mean activ­ ity of the migrant housing center families (n  894) was 14.6  2.6 nmol/min/mg Hb – virtually the same as those from fingersticks and venous blood draws of approxi­ mately 12 University of California, Davis, volunteers (Figures 68.7 and 68.8). One problem to circumvent is the possible contamina­ tion of the sample with pesticide on the skin (Yuknavage et al., 1997).

68.8  Standards Several companies provide AChE standards for using human and other species ChE preparations. Wilson et al. (2000, 2009) have been using an AChE standard prepared by hemolyzing washed bovine RBCs. These RBC ghosts show low variability when stored refrigerated (4°C) for 60 days or at low-temperature (75°C) freezer tempera­ tures for more than 250 days. Its latest use is by a clinical

1469

Table 68.14  Average Values of Cholinesterase Activity in Fasting Healthy Humansa Subjects

Men

Women

N

40

38

BuChE (U/l, mean  SD)

8920  2500

7490  1950

AChE (nU/RBC, mean  SD)

1210  200

1220  123

AChE (Hb) (U/g, mean  SD)

37.6  4.91

39.3  4.49

a

BuChE with butyrylthiocholine, Boehringer–Mannheim (Roche) kit; AChE with acetylthiocholine according to Ellman. Source: Sidell and Kamiskis (1975).

Table 68.15  ChE Activities of Healthy Human Subjects Mean  SEa

n

Male

4.99  0.14

72

Female

5.18  0.18

71

Pooled

5.08  0.11

143

Male

2.26  0.04

101

Female

2.46  0.09

71

Pooled

2.34  0.05

172

Enzyme AChE

BuChE

a

IU/ml whole blood; Ellman assay. Modified from Bellino et al. (1978).

laboratory (Wilson et al., 2009) conducting court-ordered monitoring of blood of orchard workers in the state of Washington.

68.9  Field kits Well-designed, reliable field kits for cholinesterase deter­ minations would be valuable for monitoring the health of those who apply pesticides and those who work in agricultural workplaces subject to pesticide spraying. Enzyme-impregnated filter paper, potentiometric sensors, and colorimetric comparisons have been used in the past (Collombel and Perrot, 1970; Dahlgren, 1983; Gamson et al., 1973; Rogers et al., 1991). A relatively new device is the EQM Test-Mate kit (Magnotti and Eberly, 1996; Magnotti et al., 1988). It uses a solid-state device and the Ellman method to measure ChE activity in a drop of whole

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1470

6

Number of subjects

5 4 3 2 1 0 1

3

5

7

9

11

13

15

17

19

21

23

25

27

29

31

AChE activity (nmol/min/mg Hb) Figure 68.7  Fingerstick AChE activities of University of California, Davis, volunteers. Whole blood hemolyzed with Triton X-100 and assayed according to Ellman modified for a multiple plate reader. Activity is 14.6  1.2 nmol/min/mg hemoglobin. N  13 (reproduced from Wilson et al., © 1998, Plenum).

200 180

Number of subjects

160 140 120 100 80 60 40 20 0

1

3

5

7 9 11 13 15 17 19 AChE activity (nmol/min/mg Hb)

21

23

25

27

29

31

Figure 68.8  Fingerstick AChE activities of migrant family center residents. Blood-filled capillaries kept on ice before returning to the laboratory. Whole blood assayed as in Figure 68.7. Activity is 14.6  2.6 nmol/min/mg hemoglobin. N  894 (reproduced from Wilson et al., © 1998, Plenum).

blood obtained by a fingerstick. Several models have been marketed. The Test-Mate has the advantages of portability, relative low cost, conveniently prepared reagents, and small sample volumes. Field and laboratory studies have been conducted with it (Keifer et al., 1996; Prall et al., 1998). Nevertheless, the current model (Model ChE) is not recom­ mended by the manufacturer for field use; the instructions advise that it be operated in the laboratory by a trained technician. One difficulty with using whole blood concerns sensitivity; blood samples are not diluted as much as they would be using a larger, more sensitive device. Under these conditions, readings may be affected by the relatively high

absorption of the Soret band of Hb at 412 nm, the optimum absorbance of DTNB, which is the Ellman colorimetric reagent. The Test-Mate models have attempted to circum­ vent this by using higher wavelengths, sacrificing sensitiv­ ity of the chromogen for a lower noise level (Wilson and Henderson, 1992). A second problem with the device is that the Test-Mate does not display the raw absorbance val­ ues of the reaction. Instead, it displays values normalized to those expected at 25°C, employing a temperature sensor and a built-in algorithm that has been criticized (Amaya et al., 1996; London et al., 1995; Oliveira et al., 2002; Wilson et al., 1998).

Chapter | 68  Cholinesterases

68.10  Regulatory matters: are ChE inhibitions adverse effects? Soon after the introduction of OP pesticides, the experi­ ences of University of California scientists led to a re­commendation that “individuals showing a 20% or more depletion from normal preexposure plasma ChE levels should discontinue participation in the work … until cho­ linesterase levels have returned to normal” (Metcalf, 1951). More than 50 years and many studies and task forces later, similar guidelines are still used. The reason for choos­ ing one specific decrease in ChE level over another as constituting a health hazard is not clear. One rationale is that because most clinical laboratories should be able to detect a 20% decrease in ChE activity, and because dose–response curves for many anti-ChEs tend to be steep, this “slippery slope” provides a realistic statistically signif­ icant difference between test groups suitable for regulatory purposes.

68.11  Blood ChEs and detection   of exposure Whether or not a decrease in ChE activity in the blood constitutes an “adverse effect” raises questions concern­ ing the short- and long-term health of individuals, popula­ tions, and their progeny that have yet to be answered to the satisfaction of most investigators. Detecting a statistically significant decrease in blood ChE levels between a puta­ tive exposed group and an unexposed group, or compared to an accepted “normal” range, is often accepted as indi­ cating that a potentially hazardous exposure to an anti-ChE chemical has occurred. Anti-ChE chemicals are not usu­ ally long-lasting within the body, and blood ChE activity can be expected to recover relatively rapidly from inhibi­ tion by an OP or CB. In the human, RBCs (and their AChE activities) are replaced at approximately 0.9% per day (a 120-day life span); plasma BuChE is replaced even more rapidly (Boyer et al., 1977).

68.12  Reactivation of inhibited AChE The discovery by Wilson and Ginsburg (1955) that oximes could displace OPs from the active site of ChEs, restor­ ing enzyme activity, provided an important treatment for OP poisonings. It also opened the door to using reactiva­ tion of inhibited enzymes to establish that exposure had occurred when reliable unexposed control data or normal ranges of activity were lacking (Hansen and Wilson, 1999; reviewed by Wilson et al., 1992b). Useful as it may be, the application of reactivation techniques is not currently per­ formed on a routine basis by clinical laboratories known to the author. Benschop and colleagues treated OP-inhibited serum ChE with potassium fluoride at pH 4 to chemically

1471

detect the presence of the inhibited OP-ChE. The fluoride ions reactivated the inhibited enzyme, converting the OP moiety into the corresponding phosphofluoridate and per­ mitting its subsequent identification. Benschop’s group applied the technique to obtain direct evidence of exposure to sarin in several victims of the release of the nerve gas in a Tokyo subway in 1986 and from an earlier incident in Matsumoto, Japan (Polhuijs et al., 1997, 1999).

68.13  Significance of blood ChEs Even if ChEs were not the targets of a multibillion-dollar, worldwide pesticide industry, they would still be of great interest to physiologists, cell biologists, and biochemists. ChEs have important roles in regulating neural transmis­ sion, and they are targets for pharmacological interven­ tions in disorders such as glaucoma, myasthenia gravis, and Alzheimer’s disease (Taylor, 1996). The widespread use of OP and CB pesticides and the dangers attendant upon their applications have resulted in ChE enzymes being used as biomarkers of both exposure and effect when assessing the risks to workers in agriculture and to consumers. ChEs have several roles. One is in the use of experimental animals – most often rats but also dogs, rabbits, and other species – to determine no-effect levels that can be extrapolated to humans. Another is in the recommendation of safe residue levels in food. A third role is to help provide a safe agricul­ tural workplace by monitoring those who could be exposed to dangerous levels of anti-ChE chemicals (i.e., mixer load­ ers and applicators). A fourth role is to decide whether poi­ soning episodes have involved ChE-inhibiting agents. One factor in favor of using ChEs for such clinical and regula­ tory ends is the ease with which they may be rapidly and inexpensively assayed compared to analyzing for the cho­ linergic chemicals. Generations of toxicologists and public officials have worked to establish ChE assays as a simple way to help provide answers to complicated questions of health effects, exposures, and risk. Today, agencies require submission of blood and brain ChE levels from experimen­ tal animals after short- and long-term experiments as part of the registration process for pesticides. Most agree with the position that statistically significant decreases in brain ChE activities, when accompanied by knowledge of the doses involved, are useful for establishing quantitative toxicity indices. However, issues such as the significance of ChE levels in specific parts of the brain and the applicability of one animal model over another are unresolved. There is continuing discussion of the significance of monitoring blood ChEs of humans and other animals. One issue is the role of no-observable-adverse-effect levels (NOAELs; the highest dose levels at which no important effect of a drug is observed) in assigning safe levels for toxic chemicals. One way to establish NOAELs is to per­ form batteries of behavioral tests under controlled laboratory

1472

conditions. Another is to measure residues on skin and clothing, urinary metabolites of agricultural workers, and fecal metabolites of laboratory and wild animals. Blood ChE levels represent standardized, relatively inexpensive measurements of a biochemical effect due to an exposure to a toxic chemical, in addition to providing evidence of the exposure (Nigg and Knaak, 2000). However, some do not agree. For example, an industry “Acute Cholinesterase Risk Assessment Work Group” published a review of RBC AChE, plasma ChE, and brain AChE activities focusing on their use in risk assessment (Carlock et al., 1999). The review drew upon the literature and a compilation of toxi­ city data from previously unpublished industry sources in which the chemicals were listed by code and category. The work group concluded in part that (1) plasma ChE should not be used in risk assessments because it is not an adverse effect; (2) RBC inhibition is not per se an adverse effect; and (3) when available, cholinergic effects or brain AChE levels should be used before RBC AChE values in setting NOAELs, and human data should take precedence over animal-derived results. In its discussion of NOAELs, the group did not take into account that much of the data were derived from studies of rats and dogs and were not cor­ rected for blood AChE levels or thiol oxidase activities of the RBCs, as discussed previously. A moderate position might be to use blood ChE values as biomarkers of exposure to anti-ChE inhibitors rather than to insist they be considered quantitative indicators of physiological effects, thus supporting their use as early warning signs and as important weight-of-evidence factors.

68.14  Direct effects Although much attention of neurotoxicologists has been directed toward understanding the nature of the inhibition of ChEs and their toxicological consequences, a number of neuroscientists have studied direct effects of anti-ChE agents on the presynaptic release of ACh (Rocha et al., 1996) and on the postsynaptic target receptors of ACh. For example, van den Beukel et al. (1998) found that micromolar levels of physostigmine, parathion, paraoxon, and phenyl saligenin cyclic phosphate blocked ACh-induced transient nicotinic inward currents in mouse N1E-115 and human neuroblastoma and locust thoracic ganglion cells. Eldefrawis and colleagues (Katz et al., 1997) demonstrated that chlorpyrifos, parathion, and their oxons bind to and desensitize the nicotinic receptor (nAChR) of Torpedo, the electric ray. Narahashi’s group (Nagata et al., 1997) used rat clonal pheochromocytoma (PC12) cells to demonstrate that neostigmine and carbaryl blocked nAChR channels. In con­ trast, Albuquerque’s group (Camara et al., 1997) found that methamidophos did not affect neurotransmitter release or act directly on rat nAChR but that choline affected the response of the receptors (Albuquerque et al., 1998). The low levels

Hayes’ Handbook of Pesticide Toxicology

at which these electrophysiological effects occur strongly suggest they should be taken into account when consider­ ing the effects of ChE-inhibiting OPs and CBs, such as the report of Burruel et al. (2000) of effects of methamidophos on sperm in mice.

68.15  Antidotes ChE inhibitions by OPs and CBs are examples of enzyme inhibitions for which there are specific antidotes. Two drugs in use are atropine and pralidoxime (2-PAM). Doses depend on the extent of exposure and species. Atropine binds to the muscarinic ACh receptor (mAChR), reduc­ ing the effectiveness of the excess ACh generated by the inhibition of AChE. It is given intravenously as required to relieve the symptoms of excess ACh in cases of pesticide poisoning. Oximes directly reactivate OP-inhibited AChEs [Wilson and Ginsburg (1955) reviewed in Wilson et al. (1992b)]. 2-PAM Cl (Protopam) is the oxime registered for use in the United States; its methanesulfonate salt (P2S) is used in Europe. Reactivation of the enzyme involves trans­ fer of the substituted phosphate or phosphonate residue from the catalytic site of the enzyme to the oxime. There has been much research on the treatment of exposure to ChE inhibitors focusing on chemical war­ fare agents (Hille et al., 1995; Raveh et al., 1996). The U.S. Department of Defense kit “Convulsant Antidote for Nerve Agents (CANA)” has 2 ml of the anticonvulsant diazepam, and the “Nerve Agent Antidote Kit (NAAK)” contains autoinjectors with 2 mg of atropine and 600 mg of pralidoxime. A third kit, “Nerve Agent Pre-treatment Sets (NAPS),” contains 30-mg tablets of pyridostigmine bro­ mide to be taken orally three times a day. A Swedish auto­ injector contains HI-6 (500 mg), an oxime not available in the United States, and atropine (2 mg). The logic behind the use of pyridostigmine bromide is that this CB AChE inhibitor will reduce the extent to which AChE becomes irreversibly inhibited by rapidly aging chemical warfare agents by virtue of its ability to temporarily occupy the catalytic site of AChE, interfering with its more perma­ nent phosphorylation by soman and other OPs (Tuovinen et al., 1999). Three 30-mg doses of pyridostigmine bro­ mide per day were given to the Allied Forces during the first Gulf War under an experimental-drug U.S. Food and Drug Administration license. This led to speculation that the interaction of this CB with other neuroactive chemicals present in the sector was a factor in the cluster of symp­ toms termed the Gulf War syndrome (Abou-Donia et al., 1996a,b; Kurt, 1998). An alternative approach to OP poi­ soning is to destroy the agent within the body. Two of the most successful treatments in experimental animals are to inject purified ChEs to bind the agents and to inject phos­ photriesterases to destroy them (Tuovinen et al., 1999; Wolfe et al., 1987).

Chapter | 68  Cholinesterases

Experimental evidence of nerve damage (in this case to chickens) was reported by Abou-Donia and co-workers after combined treatments of pyridostigmine bromide, DEET (an insect repellent), chlorpyrifos (Abou-Donia et al., 1996a), or permethrin (Abou-Donia et al., 1996b).

68.16  Risk assessment and ChEs ChE activity has been used for many years as a biomarker of exposure and effect for setting regulations for anti-ChE pesticide use and for safe levels of such pesticides in foods and in the environment. Almost 75 years of research has provided tools to qualitatively establish whether exposure has occurred to humans and other animals in the labora­ tory, the clinic, and the environment. Nevertheless, prob­ lems arise when data are used quantitatively, such as in setting NOAELs of a pesticide in food or in estimating lifetime exposure levels of a chemical warfare agent. The lack of universally acceptable standards for ChE assays and the difficulty in deciding whether ChE levels are, in and of themselves, an adverse effect are two of the diffi­ culties encountered when using ChE activities for regula­ tory purposes. However, cholinergic mechanisms became a criterion for creating a category of aggregate pesticide use in the Food Quality and Protection Act of 1996 (FQPA; Mileson et al., 1998). It created a single health-based safety standard for pesticide residues in food and removed the regulation that specified zero tolerance from pesticides that may be concentrated in processed commodities. FQPA required a “reasonable certainty” that no harm will result from aggregate exposures from chemicals with a common mode of action. Exposures from diet, including drinking water, and nonfood exposures (e.g., residential, lawn, gar­ den, indoor, institutional, and industrial uses) were all to be considered. Children received special treatment; up to 10 additional uncertainty factors could be added for them. The FQPA specified the use of sound science in making the determinations and required a focus on health-based approaches to food safety and the promotion of safer, effective pest control methods. Applying the policies delineated by the FQPA may occupy the attention of the agricultural community, gov­ ernment regulators, and toxicologists specializing in agri­ cultural chemicals for some time to come. It is safe to predict that there will continue to be important issues at each step of the risk assessment process in the assessments of hazard, dose–response, and exposure and in the charac­ terization of risk of a toxic chemical. A special problem involves determining aggregate exposures, deciding which chemicals should be included as part of a common “risk cup.” ChE-inhibiting chemicals played a leading role in this venture when a panel of dis­ tinguished toxicologists asserted that inhibition of AChE was a “common mechanism of toxicity” for OPs. They

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concluded that “OP pesticides act by a common mecha­ nism of toxicity if they inhibit phosphorylation and elicit any spectrum of cholinergic effects” (Mileson et al., 1998, p. 8). It may be too soon to state whether such a pharma­ cologically simplistic, but regulatory valuable, view will win the day. Pope (1999), one of the authors of the “com­ mon mechanism” review, concluded in a separate article that “additional macromolecular targets for some OP pesticides … may alter the cascade of events following AChE phosphorylation and thereby modify that common mechanism” (p. 16). His review analyzed the comparative toxicity of 38 OP AChE inhibitors currently in use in pesti­ cides. Demonstrated direct effects of anticholinesterases on receptors and transmitter release discussed previously, and the evidence of groups such as Albuquerque’s that choline may have an effect on ACh receptors (Albuquerque et al., 1998; Alkondon et al., 1997), suggest that there may be holes in the risk cup. Whether the leaks are large enough to lead to a modification of the “common mechanism” policy is a matter for the future.

Conclusion The specific course of research on ChEs parallels, and indeed has often been at the cutting edge of, some of the advances of the mid-20th and early 21st century in biochemistry, pharmacology, physiology, cell biology, and toxicology. Many of today’s disciplines had not yet been christened when ChEs were first recognized as special proteins intimately associated with the development and regulation of the nervous system. Although some anticho­ linergic agricultural chemicals have been replaced by other agents with different modes of action (e.g., imidacloprid) (Thyssen and Machemer, 1999), OP and CB use is likely to continue in the United States and abroad for the fore­ seeable future, if only because some countries may not be able to afford the new generations of chemicals. The rapid development of cDNA microarrays suggests that stud­ ies of the impact of anticholinergic chemicals will soon be routinely conducted on the level of responsive genes (Gupta et al., 1999). The advent of probabilistic methods of assessing risk of pesticides to human health and wildlife will create opportunities to apply sophisticated methods of determining risk from anticholinergic agents based on the population distributions of exposures and effects (Boyce, 1998). Research on the molecular bases of pharmaceuti­ cals, enzyme action, development of the nervous system, and many other basic features of living systems will ben­ efit from studies of ChEs for many years to come. Also, dismayingly, the simplicity of synthesizing and deploying anti-ChEs as weapons of war and terror is likely to be as tempting to governments and to terrorists of the 21st cen­ tury as it was to some in the 20th century. Today’s scien­ tists leave a legacy of knowledge for those who come after,

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but the gift of wisdom is not ours to bestow, much as we may wish to do so.

Acknowledgment I am grateful for the assistance of John D. Henderson in the preparation of this chapter.

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Chapter | 68  Cholinesterases

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Sket, D., Dettbarn, W.-D., Clinton, M. E., Misulis, K. E., Sketelj, J., Cucke, D., and Brizin, M. (1991). Prevention of diisopropylphospho­ rofluoridate-induced myopathy by botulinum toxin type A blockage of quantal release of acetylcholine. Acta Neuropathol. 82, 134–142. Spencer, P. S., Schaumburg, H. H., and Ludolph, A. C. (2000). “Experimental and Clinical Neurotoxicology”, 2nd ed. Oxford University Press, New York. Sperling, L. E., Steinert, G., Boutter, J., Landgraf, D., Hescheler, J., Pollet, D., and Layer, P. G. (2008). Characterisation of cholinesterase expression during murine embryonic stem cell differentiation. Chem. Biol. Interact. 175(1-3), 156–160. Sussman, J. L., Harel, M., Frolow, F., Oefner, C., Goldman, A., Toker, L., and Silman, I. (1991). Atomic structure of acetylcholinesterase from Torpedo californica: A prototypic acetylcholine-binding protein. Science 253, 872–879. Talesa, V., Grauso, M., Arpagaus, M., Giovannini, E., Romani, R., and Rosi, G. (1999). Molecular cloning and expression of a full-length cDNA encoding acetylcholinesterase in optic lobes of the squid Loligo opalescens, a new member of the cholinesterase family resistant to diisopropyl fluorophosphate. J. Neurochem. 72(3), 1250–1258. Taylor, P. (1994). The cholinesterases: From genes to proteins. Annu. Rev. Pharmacol. Toxicol. 34, 281–320. Taylor, P. (1996). Cholinesterase agents. In “Goodman and Gilman’s the Pharmacological Basis of Therapeutics” (J. G. Hardman et al, eds.), 9th ed., pp. 161–176. McGraw-Hill, New York. Taylor, P. (1999). Esterases reacting with organophosphorus compounds. Chem. Biol. Interact. 119/120, 1–620. Thomsen, T., Kewitz, H., and Pleul, O. (1988). Estimation of cholinester­ ase activity (EC 3.1.17; 3.1.1.8) in undiluted plasma and erythrocytes as a tool for measuring in vivo effects of reversible inhibitors. J. Clin. Chem. Clin. Biochem. 26, 469–475. Thomsen, T., Kewitz, H., and Pleul, O. (1989). A suitable method to monitor inhibition of cholinesterase activities in tissues as induced by reversible enzyme inhibitors. Enzyme 42, 219–224. Thyssen, J., and Machemer, L. (1999). Imidacloprid: toxicology and metabolism. In “Nicotinoid Insecticides and the Nicotinic Acetylcholine Receptor” (I. Yamamoto and J. E. Casida, eds.), pp. 213–222. Springer, Tokyo. Traina, M. E., and Serpietri, L. A. (1984). Changes in the levels and forms of rat plasma cholinesterases during chronic diisopropylphosphofluo­ ridate intoxication. Biochem. Pharmacol. 33, 645–653. Tuovinen, K., Kaliste-Korhonen, E., Raushel, F. M., and Hanninen, O. (1999). Success of pyridostigmine, physostigmine, eptastigmine and phosphotriesterase treatments in acute sarin intoxication. Toxicology 134, 169–178. U.S. Environmental Protection Agency (EPA) (1992). “Workshop on Cholinesterase Methodologies”. Office of Pesticide Programs, U.S. EPA, Washington, DC. van den Beukel, I., van Kleef, R. G., and Oortgissen, M. (1998). Differential effects of physostigmine and organophosphates on nico­ tinic receptors in neuronal cells of different species. Neurotoxicology 19, 777–787. Willig, S., Hunter, D. L., Dass, P. D., and Padilla, S. (1996). Validation of the use of 6,6-dithiodinicotinic acid as a chromogen in the Ellman method for cholinesterase determinations. Vet. Hum. Toxicol. 38, 249–253. Wills, J. H. (1972). The measurement and significance of changes in the cholinesterase activities of erythrocytes and plasma in man and ani­ mals. CRC Crit. Rev. Toxicol. 1, 153–199. Wilson, B. W. (1999). Cholinesterases. In “Clinical Chemistry of Laboratory Animals” (E. Quimby and W. Loeb, eds.), 2nd ed., pp. 430–440. Taylor & Francis, Philadelphia.

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Wilson, I. B., and Ginsburg, S. (1955). A powerful reactivator of alkyl phos­ phate-inhibited acetylcholinesterase. Biochim. Biophys. Acta 18, 168–170. Wilson, B. W., and Henderson, J. D. (1992). Blood esterase determinations as markers of exposure. Rev. Environ. Contam. Toxicol. 128, 55–69. Wilson, B. W., and Henderson, J. D. (2007). Determination of cholines­ terase in blood and tissue. Curr. Protocols Toxicol.(Suppl. 34), 1–16. Wilson, B. W., Jaeger, B., and Baetcke, K. (eds) (1992a). “Proceedings of the U.S. EPA Workshop on Cholinesterase Methodologies, Arlington, VA, Dec. 4–5, 1991”. Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. Wilson, B. W., Hooper, M. J., Hansen, M. E., and Nieberg, P. S. (1992b). Reactivation of organophosphate inhibited AChE with oximes. In “Organophosphates: Chemistry, Fate, and Effects” (J. E. Chambers and P. E. Levi, eds.), pp. 107–137. Academic Press, New York. Wilson, B. W., Padilla, S., and Henderson, J. D. (1996). Factors in stan­ dardizing automated cholinesterase assays. J. Toxicol. Environ. Health 48, 187–195. Wilson, B. W., Sanborn, J. R., O’Malley, M. A., Henderson, J. D., and Billitti, J. R. (1997). Monitoring the pesticide-exposed worker. Occup. Med. 12, 347–363. Wilson, B. W., McCurdy, S. A., Henderson, J. D., McCarthy, S. A., and Billitti, J. E. (1998). Cholinesterases and agriculture, humans, laboratory animals and wildlife. In “Structure and Function of Cholinesterases and Related Proteins” (B. P. Doctor et al, eds.), pp. 539–546. Plenum, New York. Wilson, B. W., Henderson, J. D., Bosworth, D. H., and Oliveira, G. H. (2000). “Standardization of cholinesterase measurements for moni­ toring human exposures”. Book of Abstracts. 219th ACS National Meeting, San Francisco, CA, March 26–30. Wilson, B. W., Henderson, J. D., Furman, J. L., Zeller, B. E., and Michaelson, D. (2009). Blood cholinesterases from Washington State orchard workers. Bull. Environ. Contam. Toxicol. 83(1), 59–61. Witter, R. F. (1963). Measurement of blood cholinesterase. Arch. Environ. Health, 6, 537–563. Wolfe, A. D., Rush, R. S., Doctor, B. P., Koplovitz, I., and Jones, D. (1987). Acetylcholinesterase prophylaxis against organophosphate toxicity. Fundam. Appl. Toxicol. 9, 266–270. Wright, C. I., and Sabine, J. C. (1948). Cholinesterases of human eryth­ rocyte and plasma and their inhibition by antimalarial drugs. J. Pharmacol. 93, 230–239. Wyttenbach, C. R., and Thompson, S. C. (1985). The effects of the organophosphate insecticide malathion on very young chick embryos, malformations detected by histological examination. Am. J. Anat. 174, 187–202. Xie, W., Stribley, J. A., Chatonnet, A., Wilder, P. J., Rizzino, A., McComb, R. D., Taylor, P., Hinrichs, S. H., and Lockridge, O. (2000). Postnatal developmental delay and supersensitivity to organo­ phosphate in gene-targeted mice lacking acetylcholinesterase. J. Pharmacol. Exp. Ther. 293(3), 896–902. Yamalik, N., Ozer, N., Caglayan, F., and Caglayan, G. (1990). Determination of pseudocholinesterase activity in the gingival crevic­ ular fluid, saliva, and serum from patients with juvenile periodontitis and rapidly progressive periodontitis. J. Dental Res. 69, 87–89. Yeary, R. A., Eaton, J., Gilmore, E., North, B., and Singell, J. (1993). A multiyear study of blood cholinesterase activity in urban pesticide applicators. J. Toxicol. Environ. Health 39, 11–25. Yuknavage, K. L., Fenske, R. A., Kalman, D. A., Keifer, M. C., and Furlong, C. E. (1997). Simulated dermal contamination with capillary samples and field cholinesterase biomonitoring. J. Toxicol. Environ. Health 51, 35–55. Zajicek, J. (1957). Studies on the histogenesis of blood platelets and megakaryocytes. Acta Physiol. Scand. 40(Suppl. 138), 1–32.

Chapter 69

Organophosphorus-Induced Delayed Neuropathy Marion Ehrich and Bernard S. Jortner Virginia–Maryland Regional College of Veterinary Medicine, Virginia Tech, Blacksburg, Virginia

69.1  History Neuropathy due to exposure to organophosphorus (OP) compounds was first reported in 1899, many years before this class of chemical agents was recognized for its insecticidal capabilities (Abou-Donia, 1981, 1995, 2003; Cherniack, 1986, 1988; Costa, 2008; Ecobichon, 1994; Ehrich and Jortner, 2001; Gallo and Lawryk, 1991; Johnson, 1982; Lotti, 1992; Lotti and Moretto, 2005; Moser et al., 2008). Until the 1930s, however, organophosphorus-induced delayed neuro­ pathy (OPIDN) appeared as isolated incidents and attracted little attention from the biomedical community. In the 1930s, between 4000 and 20,000 residents of the United States, which was under Prohibition at the time, were affected when a tricresyl phosphate-containing preparation was used as an alcohol substitute (Kidd and Langworthy, 1933). This product, called Ginger Jake, caused limb weakness and ataxia in exposed individuals from which they did not fully recover. Since the 1930s, other situations in which exposure to OP compounds has caused delayed neuropathy in humans have been identified. Some of these have been isolated incidents, involving intentional or accidental exposures of individuals. Others have resulted in the exposure of numerous people, such as those who consumed tricresyl phosphate-contaminated cooking oil in Europe, Africa, and Asia, those who used an adulterated abortifacient, and shoe-manufacturing workers in Italy exposed to an OP agent used as a plasticizer. These incidents put more than 10,000 people at risk for OPIDN (Abou-Donia, 1981, 2003; Abou-Donia and Lapadula, 1990; Cavalleri and Cosi, 1980; Cherniack, 1988; Ecobichon, 1994; Glynn, 2007; Lotti, 1992; Lotti and Moretto, 2005). Animals, too, can be susceptible to OPIDN following accidental exposures, and reports of OPIDN in water buffalo, cattle, horses, and sheep appear in the literature (Beck et al., 1977; El-Sebae et al., 1977; Perdrizet et al., 1985; Sanders et al., 1985). Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

Most of the victims of OPIDN have been people and animals exposed to OP agents that are used as lubricants and plasticizers rather than insecticides. Current federal testing of OP compounds under the U.S. Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA) requires that those proposed for use as insecticides be examined for their capability to cause OPIDN in relation to their capability to inhibit acetylcholinesterase, which is their insecticidal mechanism of action [U.S. Environmental Protection Agency (EPA), 1996]. Consequently, unless exposure is at concentrations that are above those necessary to cause signs of acute toxicity due to acetylcholinesterase inhibition, humans and susceptible animals are unlikely to develop OPIDN following insecticide exposure. OPIDN has been suggested to be involved in other disorders, such as that seen in veterans returning from the 1991 Gulf War (Haley et al., 1997). However, neurological examination of these veterans and animal studies do not support that association (Institute of Medicine, 2003, 2004). Soldiers on active duty were exposed to a variety of chemicals, including OP insecticides, pyrethrin insecticides, insect repellents, petroleum products, and sand dust. They were also at risk for exposure to OP compounds used as chemical warfare agents (potent acetylcholinesterase inhibitors), and to prevent toxicity associated with such, the carbamate acetylcholinesterase inhibitor pyridostigmine was used to protect them preexposure. Soldiers were vaccinated against infectious diseases, anthrax, and botulinum toxin (Institute of Medicine, 2001). Investigations have determined that OP chemical warfare agents were released during this conflict; exposure levels were not determined, but levels were too low to cause immediate physical symptoms. Although OPIDN appears weeks to months after exposure to some OP compounds, the OP insecticides used in the Persian Gulf and the OP chemical warfare agents are unlikely to cause this syndrome, especially in the absence 1479

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of clinical signs of acute acetylcholinesterase inhibition (Institute of Medicine, 2003, 2004).

69.2  Chemistry of organophosphorus compounds Not all OP compounds are capable of causing OPIDN. Lists of neuropathy-inducing OP compounds with pentavalent phosphorus ions are contained in the 1991 edition of the Handbook of Pesticide Toxicology (Gallo and Lawryk, 1991) and other reviews of OPIDN (Abou-Donia, 1981; Abou-Donia and Lapadula, 1990; Cherniack, 1988; Hollingshaus, 1983; Johnson, 1975, 1982). Among the OP compounds responsible for reported incidents of OPIDN in man and animals or OP compounds used for laboratory studies of OPIDN are the protoxicants leptophos (O-4-bromo-2,5-dichlorophenyl O-methyl phenylphosphorothioate) and tri-ortho-cresyl phosphate [TOCP; also known as tri-ortho-tolyl phosphate (TOTP)], which is metabolized to a cyclic saligenin phosphate responsible for inducing OPIDN. Active OPIDNinducing agents used for many experimental studies include mipafox (N,N’-diisopropylphosphorodiamidic fluoride), diisopropyl phosphorofluoridate (DFP), and di-n-dibutyl2,2-dichlorovinyl phosphate (DBDCVP) (Figure 69.1). None of these are currently used as insecticides. Several reports and reviews, including those listed previously and others (Abou-Donia, 1995; Ecobichon, 1994; Johnson and Glynn, 2001; Johnson et al., 1991; Kropp and Richardson, 2003; Lotti, 1992; Malygin et al., 2003; Wu and Casida, 1994; Yoshida et al., 1994), have examined structure–activity relationships among compounds that cause OPIDN, with the following points noted: 1. For the classical OPIDN discussed in this chapter, phosphorus must be in the pentavalent state. 2. The atom with the coordinate covalent bond attached to the phosphorus must be an oxygen; protoxicants with sulfur attached by a coordinate covalent bond can be oxidized to active neurotoxicants. 3. The neuropathy-inducing OP compounds all have at least one oxygen or amine bridge linking an R group to phosphorus. Therefore, the major subgroups producing OPIDN are phosphates (derivatives of phosphoric acid, which has four oxygens on the phosphorus), phosphonates (derivatives of phosphonic acid, which has three oxygens on the phosphorus), phosphoramidates (derivatives of phosphoramidic or phosphorodiamidic acids, with one or two nitrogens and two or three oxygens on the phosphorus), or phosphorofluoridates (three oxygens and a fluoride on the phosphorus) (Figure 69.2). 4. Alkyl substitution on an ortho site of phenyl phosphates increases the likelihood that the compound can be metabolized to a neurotoxicant. Ortho methyl substitution rather than longer chain substitution on the phenyl

CH3

O O

P

CH3

O

O O O

CH3

P O O

Phenyl saligenin cyclic phosphate (PSP)

TOCP

CH3 H3C

CH NH O P NH F

Cl

O P

O

OCH3

Br Cl

Leptophos

H 3C

CH CH3

Mipafox

CH3 H3C

CH O

O P

O H 3C

CH

CH3CH2CH2CH2O F

CH3CH2CH2CH2O

O P

OCH

CCl2

CH3

DFP

DBDCVP

Figure 69.1  Chemical structures of organophosphorus compounds commonly used in laboratory studies. Included are tri-ortho-cresyl phosphate [TOCP; also known as tri-ortho-tolyl phosphate (TOTP)]; PSP, a cyclic saligenin phosphate metabolite congener of TOCP; leptophos; mipafox; diisopropyl phosphorofluoridate (DFP); and di-n-dibutyl-2,2dichlorovinyl phosphate (DBDCVP).

ring(s) of triphenyl phosphates increases the capability to cause OPIDN. However, not all ortho-methylsubstituted phenyl phosphates induce neuro­pathy. Substitution at other sites on the ortho-substituted ring decreases the neurotoxicity. 5. Increasing the size of the alkyl substituents (up to four or five) on phosphoro- and phosphonofluoridates, phosphorodiamidofluoridates, and dichlorovinyl phosphates increases the hydrophobicity and increases the capability to cause OPIDN. 6. Chirality can contribute to neurotoxicity. Racemic mixtures tend to be less potent as inducers of OPIDN; for compounds tested to date, one enantiomer generally appears to be a more potent neuropathy-inducing agent than the other (Battershill et al., 2004; Hollingshaus, 1983; Johnson and Read, 1987; Johnson et al., 1991; Lotti et al., 1995).

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R

O

Phosphates R

69.3  Clinical manifestations

O P

O

X

O

RH2C Phosphonates

O P

O

X

O

X

RH2CO

RHN Phosphoramidates

O P

R

O

RHN Phosphorodiamidates

O P

O

R

O

Phosphorofluoridates

O

X

O

X

O

X

O P

RH2C Phosphinates

O P

69.3.1  Human Descriptions of the clinical manifestations of OPIDN in humans can be found in several reviews (Abou-Donia, 1995; Cherniack, 1986; Ecobichon, 1994; Gallo and Lawryk, 1991; Lotti, 2001; Lotti and Moretto, 2005). These descriptions follow a similar scenario. Some days (usually 6–14) after exposure, humans note tingling of the hands and feet, followed by sensory loss in the hands and feet. Electromyograms and nerve conduction studies indicate decreased firing of motor units and slowed motor conduction. However, it is the appearance days to weeks after exposure of bilateral and symmetrical weakness progressing to flaccidity of the distal skeletal muscles of the lower and upper extremities that is characteristic of this disorder. Ataxia can be noted. Although this may resolve with time, victims of OPIDN may still have abnormal reflexes and spasticity.

69.3.2  Animal

RHN

R

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RH2C Figure 69.2  Basic structures of organophosphorus compounds. Phosphates, phosphonates, phosphoramidates, phosphorodiamidates, and phosphorofluoridates cause classical OPIDN; phosphinates do not.

Neuropathy-inducing OP compounds have a leaving group attached to a labile oxygen or nitrogen bond (Johnson, 1982; Johnson and Glynn, 2001). Dealkylation at this site results in a negatively charged phosphoryl group, and formation of this chemical species is needed if a phosphate or phosphonate is to induce classical delayed neuropathy. Deprotonation rather than dealkylation is thought to provide the necessary negative charge for neuropathy-inducing phosphoramidates (Kropp et al., 2004; Richardson, 1995). This dealkylation occurs after the OP inhibits an esterase, specifically neuropathy target esterase (NTE; also known as neurotoxic esterase). The NTE inhibition and the dealkylation reaction render the structure capable of essentially irreversibly binding to NTE, which is a necessary prerequisite to OPIDN.

The adult hen is the recognized animal model for OPIDN (U.S. EPA, 1996). This is because clinical signs, which occur after a delay period similar to that which occurs in humans, are easy to observe as they progress, and the associated histopathologic changes are easily identified. In addition, the hen provides readily reproducible results in a relatively small, relatively accessible animal model. In the hen, a lag period of several days is needed before any clinical alterations appear. Early signs of OPIDN in this species are abnormal foot placement and leg weakness, which may affect balance. As OPIDN progresses, hens become reluctant to walk, show incoordination when they do, and may have wing droop and/or use wings for balance. More severe signs include loss of upright posture when walking, followed by loss of ability to walk. Eventually, the wings also become involved. These clinical deficits can be differentiated with grading systems that designate effects from mild to paralysis, such as using scoring systems that range from 1 to 5 or from 1 to 8 (Figure 69.3). However, it is only the adult chicken that shows these clinical signs upon exposure to neuropathy-inducing OP compounds. This manifestation of OPIDN does not seem to appear in chickens less than 55–60 days of age (Funk et al., 1994a; Moretto et al., 1991; Peraica et al., 1993). Clinical manifestations of OPIDN are seen in a variety of other species (Abou-Donia, 1981; Johnson, 1982), including other avians such as the pheasant and turkey (Johnson, 1982; Larsen et al., 1986). Humans and mammals such as sheep, pigs, cattle, water buffalo, horses, cats, dogs, and ferrets are susceptible to clinical manifestations of OPIDN, as indicated by progressive ataxia (Abou-Donia, 1981, 1995; Johnson, 1982; Jortner et al., 1983; Stumpf et al., 1989). Clinical manifestations specifically associated with OPIDN are not,

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69.4  Neuropathology

Clinical score

4

3

2

1 10

12

14 18 Days after organophosphate

21

Figure 69.3  Development of clinical signs in chickens after administration of phenyl saligenin phosphate (PSP) and tri-ortho-tolyl phosphate (TOTP). Results are presented as the mean  SD, n  3–9, on a scale of 1–5. PSP im 2 mg/kg (—), 3 mg/kg (—), and 10 mg/kg (—); TOTP 360 mg/kg po (—) and 500 mg po (—). Increasing clinical scores reflect progression of deficits. Reproduced with permission from Jortner, B. S., and Ehrich, M. (1987). Neuropathological effects of phenyl saligenin phosphate in chickens. Neurotoxicology 8, 303-314. Copyright © 1987 by Intox Press, Little Rock, AR.

however, generally seen in laboratory rodents (Ehrich et al., 1995, 2004; Lehning et al., 1996; Padilla and Veronesi, 1985; Somkuti et al., 1988; Veronesi et al., 1991), although they have been reported to appear in rats more than 6 months old or mice dosed for more than 200 days (Lapadula et al., 1985; Moretto et al., 1992b). OPIDN is induced by pentavalent phosphorus compounds (phosphates, phosphonates, and phosphoramidates). However, a different type of delayed neuropathy can be induced by phosphites (trivalent phosphorus compounds), with most studies done with triphenyl phosphite. The latter has been labeled type II OPIDN (Abou-Donia and Lapadula, 1990). Triphenyl phosphite has been reported to cause ataxia in species both susceptible to (hens and ferrets) and relatively resistant to (rats and Japanese quail) classical OPIDN. The latent period for type II OPIDN is short, and prominent clinical manifestations, such as hyperexcitability, spasti­city, tail kinking, side-to-side movements, circling, ataxia, and flaccid paralysis, occur in rodents (Lehning et al., 1996; Veronesi and Dvergsten, 1987). Pathological manifestations of phosphite neurotoxicity also differ from those characteristic of OPIDN (Abou-Donia and Lapadula, 1990; Lehning et al., 1996). Some investigators believe that based on differences in species susceptibility, magnitude of NTE inhibition, onset and nature of clinical signs, and extent and nature of brain and spinal cord lesions, these are not two types of OPIDN but, rather, represent separate categories of neurotoxicities induced by OP compounds (Lehning et al., 1996).

Neuropathologic studies of experimental OPIDN induced by pentavalent phosphorus compounds have revealed a consistent pattern of lesions, which is believed to represent the morphologic substrate of the entity. The test compounds most often used to elicit these lesions were TOCP or TOTP, its neurotoxic cyclic congener phenyl saligenin phosphate (PSP), DFP, or mipafox, with the chicken and cat being the major experimental animal subjects (Bischoff, 1967, 1970; Cavanagh, 1954, 1964; Cavanagh and Patangia, 1965; Ehrich and Jortner, 2001; Illis et al., 1966; Itoh et al., 1984, 1985; Jortner and Ehrich, 1987; Krinke et al., 1979; Prineas, 1969; Tanaka and Bursian, 1989). This body of work reveals that the primary lesion is a bilateral degenerative change in distal levels of axons and their terminals, primarily affecting larger/ longer myelinated central and peripheral nerve fibers, leading to breakdown of affected neuritic segments and secondarily of their myelin sheaths. These lesions generally begin to develop at or near the end of the postdosing symptom-free period. In experimental studies of chickens, this distal pattern of injury is manifest by clinical neuropathy associated with bilateral central nervous system long-tract involvement, such as in cervical spinal cord, medullary, and cerebellar levels of the ascending spinocerebellar tracts and fasciculus gracilis, and lumbar levels of the descending medial pontine spinal tracts (AbouDonia and Preissig, 1976a,b; Cavanagh, 1954; Classen et al., 1996; Itoh et al., 1984; Jortner and Ehrich, 1987; Tanaka and Bursian, 1989; Tanaka et al., 1990) (Figure 69.4). The most sensitive histological indicator of OPIDN was believed to be degenerating cerebellar fibers, especially in folia IV and V (Classen et al., 1996). Use of Fink–Heimer silver impregnation histological techniques in hens dosed with TOCP or DFP revealed more extensive distribution of central nervous system axonal and terminal degeneration, particularly extension of alterations of the lumbar medial pontine spinal tract into ventral gray matter laminae VI and VII and those of the rostral spinocerebellar system, which were seen in the deep cerebellar nuclei and mossy fiber projections to the anterior lobe of the cerebellar cortex (Tanaka and Bursian, 1989; Tanaka et al., 1990). Terminal and preterminal degeneration was also noted in a number of medullary structures, such as the spinal lemniscus and lateral vestibular, inferior olivary, gracile, external cuneate, and lateral cervical nuclei. Degenerating presynaptic boutons and small axons (Bischoff, 1970; Dyer et al., 1996) are the likely ultrastructural substrate of these silver-impregnated altered gray matter neurites in hens. In addition to these central nervous system lesions of OPIDN, distal regions of long peripheral nerve myelinated fibers in chickens, particularly in the legs, are similarly affected (Cavanagh, 1954; Dyer et al., 1991, 1992; El-Fawal et al., 1988, 1990b,c; Jortner, 1984; Jortner and Ehrich, 1987; Prineas, 1969). Attention has been directed to the sensitivity

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Figure 69.4  Cross section of a spinocerebellar tract in the cervical level of spinal cord from a hen 21 days after exposure to a neurotoxic dose of mipafox (30 mg/kg ip). The section shows extensive myelinated fiber degeneration, manifest by pale-staining swollen axons or dark-staining fibers with collapsed axons and disordered myelin sheaths. Toluidine blue–safranin stain; scale bar  25 m.

of the nerve supplying the biventer cervicis muscle and the large-diameter myelinated fibers of the tibial nerve branch to the lateral head of the gastrocnemius muscle in this species (El-Fawal et al., 1988, 1990c; Krinke et al., 1979). A similar neuropathologic pattern has been noted in cats exposed to neurotoxic doses of DFP or TOCP. This is manifest by distal degeneration mainly affecting large, long myelinated fibers, involving rostral (cervical, medullary, and, for some, cerebellar) levels of ascending tracts (fasciculus gracilis, spinocerebellar tracts, and spino-olivary tract) and thoracolumbar spinal cord levels of descending tracts (corticospinal, reticulospinal, and rubrospinal) (Abou-Donia et al., 1986; Cavanagh and Patangia, 1965). Degenerating nerve fiber terminals were seen in gray matter afferent nuclei (Cavanagh and Patangia, 1965; Illis et al., 1966; Prineas, 1969). Distal levels of long peripheral nerve myelinated fibers and their terminals, such as in the hind legs and in recurrent laryngeal nerves, were similarly affected (Bouldin and Cavanagh, 1979a,b; Cavanagh, 1964; Drakontides et al., 1982; Glazer et al., 1978; Prineas, 1969). By light microscopy, the qualitative changes were best demonstrated in sections from epoxy resin-embedded preparations of spinal cord and medullary white ­matter and peripheral nerve. These included axonal swelling with attenuated myelin sheaths (Figure 69.5). The swollen axons were pale staining or debris laden and darker staining (Figures 69.5 and 69.6). Another feature, common in central nervous system myelinated tracts, was the presence of contracted darker-staining axons with disordered myelin sheaths (see

Figures 69.4 and 69.5). These lesions increased in affected regions as the clinical neuropathy advanced. Later lesions included fragmentation of affected fiber segments as the process of Wallerian-like degeneration ensued (see Figure 69.6). This was associated with formation of myelin-rich ovoids and phagocytosis of the degraded element by macro­ phages or Schwann cells (the latter in peripheral nerve) (see Figure 69.6). Ultrastructurally, changes in axonal membrane systems were early morphologic events in OPIDN. This includes proliferation of intra-axonal tubules and cisterns, resembling smooth endoplasmic reticulum, and vesicles in hens and cats receiving toxic doses of TOCP or DFP (Bischoff, 1967, 1970; Bouldin and Cavanagh, 1979a,b; LeVay et al., 1971; Prineas, 1969) (Figure 69.7). Other early changes, seen in cats dosed with DFP, were peripheral myelinated fiber distal, nonterminal varicosities due to the presence of abnormal membrane-lined vacuoles in axons, inner myelin sheaths, or both, which preceded fiber degeneration (Bouldin and Cavanagh, 1979a,b). These workers suggest such vacuolar alterations represent a “chemical transection” of the fiber, leading to its subsequent breakdown. Following these early ultrastructural events, a variety of subsequent degradative axonal changes are seen, progressing to fiber degeneration. One sequence involves axonal swelling with secondary attenuation of the myelin sheath. Electron microscopic study revealed that numbers of these swollen axons contained disorganized masses of normal and altered mitochondria, cytoskeletal components (neurofilaments),

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Figure 69.5  High-power view of spinal cord long-tract degeneration in cross section from a chicken given mipafox (30 mg/kg ip) 21 days earlier shows a large central swollen axon with moderately dense staining axoplasm containing particulate material and a vacuole. The myelin sheath is thin. Many of the adjacent degenerating fibers have contracted, dark-staining axons with disordered dark-staining myelin sheaths (arrowhead). Pale-staining regions of myelinated fiber loss with associated gliosis are present (asterisk). Toluidine blue–safranin stain; scale bar  10 m.

Figure 69.6  This tangentially sectioned tibial nerve branch to the lateral head of the gastrocnemius muscle from a chicken 15 days after dosing with 2.5 mg/kg of phenyl saligenin phosphate shows several stages of myelinated nerve fiber degeneration (arrows). These include swollen axons with pale or moderate staining of their contents and formation of myelin-rich segments (ovoids) of Wallerian-like degeneration. Toluidine blue–safranin stain; scale bar  25 m.

lysosome-like dense bodies, and membranous multilamellar bodies (Bischoff, 1967, 1970; Bouldin and Cavanagh, 1979b; Ehrich and Jortner, 2001; Jortner and Ehrich, 1987; Prineas, 1969) (Figure 69.8). A second appearance of the swollen

axons, which may be derived from the preceding, is one in which there has been granular degeneration of its contents due to lysis of the cytoskeleton and other axonal contents, leading to swollen electron-lucent axons (Jortner and Ehrich,

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Figure 69.7  Electron micrograph showing early ultrastructural axonal changes of OPIDN (hen given PSP 2.5 mg/kg im). An increase in membranelined tubules and cisterns suggestive of proliferation of agranular endoplasmic reticulum is noted (straight arrow). In addition, there is a focus of granular degeneration of the cytoskeleton (arrowhead) and a small lamellar body (curved arrow). Scale bar  1 m.

Figure 69.8  Electron micrograph demonstrating extensive myelinated fiber degeneration in a cross section of the medullary (distal) level of the fasciculus gracilis from a chicken administered a toxic dose (500 mg/kg po) of tri-ortho-tolyl-phosphate 21 days earlier. The morphologic presentations of fiber degeneration include swollen axons with multilamellar membranous aggregates (arrow) and contracted electron-dense axons with disordered myelin sheaths (arrowhead). Scale bar  5 m.

1987) (Figure 69.9). This is thought to be an advanced manifestation of the neuropathy (Prineas, 1969) and may be related to increased activity of calcium-activated proteinases, associated with toxicant-induced intra-axonal elevations of calcium ions (El-Fawal et al., 1990a). Yet a third ultrastructural appearance of degenerating fibers, particularly prominent in the central nervous system, is distal axonal collapse with ill-defined electron-dense axoplasm and disordered myelin sheaths (see Figure 69.8). These axonal changes are associated with aggregation of membranous masses, altered mitochondria, and dense bodies in degenerating axon terminals (Drakontides et al., 1982; Glazer et al., 1978; Prineas, 1969). Some workers suggest that axon terminals are the initial site of degeneration, with subsequent involvement of terminal portions of the axon, creating a true “dying-back” neuropathy (Tanaka and Bursian, 1989). This contrasts with the view that the terminal lesions are secondary to injury in the distal, nonterminal portions of the axon (Bouldin and Cavanagh, 1979a,b; Prineas, 1969). Eventually, the degrading segments of affected myelinated fibers are phagocytized, an event occurring more rapidly in peripheral than in central regions of the nervous system (Figure 69.10). In the former, most fiber debris was phagocytized and degraded by 4 weeks following a single toxic dose in hens, but these damaged neurites may persist much longer in the central nervous system (Jortner et al., 1989). In peripheral nerve, the damaged fiber swells, fragments, and is

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Figure 69.9  Electron micrograph demonstrating one pattern of advanced axonal alteration. This is manifest by extensive granular degeneration of the axoplasm and axonal organelles in this cross-sectioned myelinated fiber from the dorsal metatarsal nerve of a chicken dosed with 10 mg/kg of phenyl saligenin phosphate 14 days earlier. Scale bar  1 m.

p­ hagocytized by macrophages or Schwann cells within the column formed by the original Schwann cell basal lamina. This resembles events in Wallerian (hence the term Wallerianlike) degeneration. With advanced breakdown of the fiber in OPIDN, proliferation of Schwann cells in their basal lamina forms the band of Büngner. The latter is a site of subsequent nerve fiber regeneration, which is robust in OPIDN following a single dose of the toxicant (Jortner et al., 1989). The bands of Büngner (columns of proliferating Schwann cells) provide an appropriate structural and growth-enhancing environment to permit re-innervation to occur. This regeneration included replacement of degenerated peripheral axon terminals as well (Glazer et al., 1978; Illis et al., 1966). Consistent with other forms of nerve fiber degeneration, there is a failure of such axonal regeneration in the central nervous system in OPIDN (Jortner et al., 1989). The prominence of spinal cord lesions relative to those in sciatic nerve of hens sacrificed 1 month or more following dosing with OPIDN-inducing insecticides was attributed to peripheral nerve regeneration (Abou-Donia and Graham, 1978; AbouDonia and Preissig, 1976a,b; Abou-Donia et al., 1979). In the central nervous system, macrophages provide the phagocytic element acting on degraded myelinated fibers, and there is prominent astrocytic proliferation in the damaged levels of spinal cord and brainstem (see Figure 69.10) (Jortner et al., 1989). Fluoro-Jade staining has been used to detect damage to brain and spinal cord of hens exposed to neuropathy-inducing OP compounds (Carlson and Ehrich, 2004). Although

sensitive, it does not detail the organelle damage seen with the preparations described previously. In contrast to the foregoing, a somewhat different pattern of nervous system lesions is seen in animals exposed to trivalent aryl organophosphates, exemplified by triphenyl phosphite (Abou-Donia and Lapadula, 1990). This entity (the so-called type II OPIDN) has been induced in rats, mice, ferrets, chickens, and Japanese quail (Carrington et al., 1988; Ehrich and Jortner, 2001; Lehning et al., 1996; Stumpf et al., 1989; Tanaka et al., 1992; Varghese et al., 1995; Veronesi and Dvergsten, 1987). One striking difference from the OPIDN induced by pentavalent OP compounds described previously is the prominent degeneration of neuronal cell bodies in regions such as the spinal cord, dorsal root ganglia, and brainstem. As in OPIDN induced by pentavalent phosphorus-containing OP compounds, bilateral degeneration of myelinated axons and terminals is prominent. However, this process is more widespread in neuropathology induced by triphenyl phosphate, with lesions being noted in peripheral nerve, spinal cord gray and white matter, brainstem nuclei and fiber tracts, cerebellum, basal ganglia, and cerebral cortex.

69.5  Neuropathology of mammalian animal models The most reliable experimental animal model of OPIDN is obtained by single or multiple dosing in the domestic adult

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Figure 69.10  This electron micrograph shows a cross-sectioned myelinated fiber in the medullary level of the fasciculus gracilis from a chicken 21 days after a neurotoxic dose of tri-ortho-tolyl-phosphate (500 mg/kg po). The degenerated axonal contents have been phagocytized by a macrophage (arrow) within the myelin tube. Scale bar  2 m.

Figure 69.11  Cross sections of sural nerve from a control rat (A) and one exposed to multiple neurotoxic doses of TOTP, plus chlorpyrifos and corticosterone (B). The latter shows the axonopathy progressing to degeneration of myelinated fibers of OPIDN (arrows), which was related to TOTP exposure (Jortner et al., 2005). Toluidine blue–safranin stain. Scale bars  20 m.

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chicken (hen), and the spectrum of nervous system alterations in that species has been documented previously in some detail. A good deal of earlier work employed the domestic cat, which, along with other susceptible mammalian species (sheep, cattle, nonhuman primates, etc.), has a qualitative and distributive pattern of lesions that follows that of the hen (Abou-Donia, 1981; Cavanagh and Patangia, 1965; Jortner, 1984, 1988; Jortner et al., 1983; Prineas, 1969). Thus, the basic nature of changes in these mammalian species has already been considered. Due to concern about regulatory reliance on an avian experimental model, there has been considerable interest in the evaluation of laboratory rodents as potential model systems for OPIDN. Most of this attention has focused on the laboratory rat (Jortner, 1988). The rat has been used in several experimental studies of OPIDN. In one, male Long–Evans rats were given multiple exposures to the organophosphates chlorpyrifos and TOTP over a 63-day period, followed by a 4-week recovery period (Jortner et al., 2005). There was concurrent administration of corticosterone to mimic the effects of chronic stress. Features of OPIDN were largely related to TOTP dosing. These consisted of brain NTE inhibition and bilateral degeneration of long myelinated fibers in peripheral nerves and spinomedullary gracile fasciculus, more severe

Hayes’ Handbook of Pesticide Toxicology

distally (Figures 69.11 and 69.12). The primary lesion was axonal degeneration progressing to fiber degeneration, with severity progressing in the 4-week recovery period. Advanced lesions in the gracile fasciculus also contained dystrophic axons (Figures 69.12, 69.13, and 69.14). The latter have been previously noted in rat OPIDN (Veronesi et al., 1986). This study demonstrated the need for multiple exposures to neurotoxic organophosphates such as TOTP to elicit prominent lesions of OPIDN (Ehrich et al., 2004; Jortner et al., 2005). Other pathological studies of rats with OPIDN were conducted using a variety of strains (Wistar, Long–Evans, and Sprague–Dawley) and single doses of TOCP (TOTP) or mipafox as toxicants (Carboni et al., 1992; Dyer et al., 1992; Ehrich et al., 1995; Inui et al., 1993; Lehning et al., 1996). The lesions were primarily found in the medullary and cervical (distal) levels of the fasciculus gracilis and its afferent target nucleus, and axonal vacuolization and swelling were prominent features (Carboni et al., 1992; Dyer et al., 1992) (Figure 69.15). Studies by Veronesi and colleagues using mipafox in rats demonstrated a similar distribution of lesions (Veronesi et al., 1986). In a series of studies in the Long–Evans strain of rat, using several dosing paradigms of TOCP, bilateral distal

Figure 69.12  Cross section of the distal (medullary) level of gracile fasciculi (arrows) showing extensive bilateral myelinated fiber degeneration of OPIDN (A). Higher power image (B) better demonstrates the presence of TOTP-induced degenerating myelinated fibers manifest by disordered masses of myelin (arrowheads) and large dystrophic axons (arrows). Rat exposed to multiple neurotoxic doses of TOTP plus corticosterone over a 63-day period. Toluidine blue–safranin stain.

Chapter | 69  Organophosphorus-Induced Delayed Neuropathy

myelinated fiber axonopathy progressing to fiber degeneration was demonstrated (Padilla and Veronesi, 1985; Veronesi, 1984). These involved distal levels of long ascending and descending spinal cord tracts and peripheral nerve, which largely recapitulated the distribution of lesions seen in hens (Cavanagh, 1954; Jortner and Ehrich, 1987). The fiber lesions developed after a postdosing latent period and were associated with a transient postdosing inhibition of wholebrain NTE. No definitive clinical deficits were observed in affected rats. Rats generally required higher dosages of toxicant to elicit pathological changes than did chickens, which sometimes created problems with acute neurotoxi­city from compounds inhibiting acetylcholinesterase (Ehrich et al., 1995, 2004). These findings – clinical insensitivity and restricted lesions – limit the utility of the laboratory rat as a model for OPIDN. The ferret is another experimental animal that has been used to study OPIDN. Studies with TOTP and DFP demonstrated toxicant-induced inhibition of brain NTE, neurologic signs, and bilateral distal/terminal axonal degeneration in the spinal cord, brainstem, and cerebellum (Stumpf et al., 1989; Tanaka et al., 1991). These studies

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indicate that neurotoxic responses of the ferret make it an appropriate model of OPIDN, although it has other limitations (availability and husbandry concerns) regarding its use as a test animal for safety assessment. Neuropathologic studies of a putative OPIDN model in laboratory mice have been done in animals given doses of TOCP higher than needed to elicit OPIDN in hens (Veronesi et al., 1991). Lesion distribution was not consistent with that seen in other models of the neuropathy, although myelinated fibers had qualitatively the same axonal pathology. The latter included intra-axonal vacuoles, neurofilament masses, and floccular degeneration, which were noted in cervical spinal cord white matter, medullary inferior olivary nucleus, and the fasciculus gracilis. The lesions were not associated with a specific pattern of toxicant-induced inhibition of NTE. Wu and Casida (1996) did elicit inhibition of neurotoxic esterase in mice exposed to either 2-octyl-4H-1,3,2-benzodioxaphosphorin-2-oxide (OBDPO) or ethyl octylphosphonofluoridate (EOPF). However, the lesions described – neuronal necrosis and edema of cerebral hemispheres, hippocampus, and brainstem – are inconsistent with those of OPIDN. Thus, the mouse is not considered to be an acceptable model of OPIDN.

Figure 69.13  Electron micrograph of a cross-sectioned dystrophic axon in the gracile fasciculus from a rat given multiple neurotoxic doses of TOTP. The enlarged axon has prominent tubulovesicular masses (arrow).

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Figure 69.14  Higher power view of a region of a dystrophic axon in the gracile fasciculus showing tubulovesicular profiles (left) and adjacent abnormal mitochondrial aggregates.

Because the genome of the mouse is frequently manipulated, murine models with disruption of the NTE gene have been created. Total loss of NTE resulted in embryonic death. Heterozygous NTE mutant mice demonstrated increased motor activity, but exposure to a potent OP inhibitor of NTE (EOPF) did not result in neuropathological effects in either the mutant or the wild-type mice (Winrow et al., 2003). Mice with brain-specific deletion of NTE demonstrated vacuolation in hippocampal and thalamic neurons even without exposure to neuropathy-inducing OP compounds (Akassoglou et al., 2004).

69.6  Pathogenesis Factors important in the development of OPIDN have been discussed (Abou-Donia, 1981, 1995; Abou-Donia and Lapadula, 1990; Cherniack, 1986; Ecobichon, 1994; Ehrich, 1996; Ehrich and Jortner, 2001; Gallo and Lawryk, 1991; Glynn, 2006, 2007; Hollingshaus, 1983; Johnson, 1982, 1992; Johnson and Glynn, 2001; Lotti, 1992; Lotti and Morello, 2005; Lotti et al., 1993; Richardson, 1995). A number of events that occur during the development of the

neuropathy have been identified, yet the temporal sequence of the precise mechanism(s) responsible for OPIDN remains elusive. One factor known to be important in the initiation of OPIDN is inhibition of NTE (also known as neurotoxic esterase). This enzyme is a molecular target of neuropathyinducing OP compounds with a pentavalent phosphorus atom, but NTE may be more of a biomarker rather than the single, specific target that initiates OPIDN (Ehrich, 1996; Ehrich and Jortner, 2001; Johnson and Glynn, 2001; Lotti and Moretto, 2005). What is certain is that NTE must be phosphorylated and extensively inhibited by the OP compound before notable OPIDN develops. Although dose– response effects have been reported both in hens and in rats (Ehrich et al., 1995), it is usually expected that NTE inhibition greater than 70% within 24–48 h after exposure is most likely to result in frank OPIDN. In addition to inhibiting NTE, the OP compounds inducing delayed neuropathy must bind sufficiently strongly to the NTE so that it is difficult or impossible to remove them (Johnson, 1982; Johnson and Glynn, 2001; Kropp et al., 2004; Richardson, 1995). For this type of binding to occur, the pentavalent OP compound

Chapter | 69  Organophosphorus-Induced Delayed Neuropathy

Figure 69.15  Medullary (terminal) level of the fasciculus gracilis from a Long–Evans rat that had been given a toxic dose (30 mg/kg ip) of mipafox 21 days earlier. There are numerous pale-staining swollen axons seen bilaterally (A). Higher power of the lesion (B) shows associated thin myelin sheaths. Toluidine blue-safranin stain.

(phosphate, phosphonate, and phosphoramidate) has a leaving group whose removal or rearrangement results in a negatively charged moiety on the enzyme (Johnson, 1982, 1992; Kropp et al., 2004). This process has been called “aging.” The correlation between early NTE inhibition and OPIDN is sufficiently strong that registration of OP insecticides under FIFRA requires that data on NTE inhibition and on doses required for its inhibition relative to acetylcholinesterase inhibition be obtained as a biochemical determinant of potential to cause OPIDN (U.S. EPA, 1996). Although early and significant inhibition of NTE is an excellent predictor of potential for developing OPIDN, the relationship between NTE inhibition and OPIDN is less clear. Items of discussion include the following: 1. Although NTE is inhibited soon after OP administration, it may no longer be inhibited some days to weeks

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later when clinical and pathological manifestations of OPIDN are evident. If NTE were a target, its inhibition would need to initiate some series of events that no longer required its inhibition (Ehrich, 1996; Johnson and Glynn, 2001; Lotti et al., 1993). 2. NTE inhibition predictive for OPIDN is usually measured in brain and spinal cord – tissues that have a relatively low proportion of their nerve fiber population affected. The most concentrated nerve fiber lesions of OPIDN are found in distal levels of peripheral nerves, such as branches of the tibial or in the biventer cervicis in the hen, and these tissues are too small for measurement of NTE activity (El-Fawal et al., 1988, 1990c; Jortner and Ehrich, 1987). The sciatic nerve has been used for assay, even though it does not show exceptional pathogenic effects, except in its distal branches, following administration of neuropathy-inducing OP compounds (Dyer et al., 1991). Furthermore, OP-induced inhibition of NTE may also occur in nonnervous tissue, including the adrenal gland, lymphocytes, and platelets (Bertoncin et al., 1985; Ehrich, 1996; Maroni and Bleecker, 1986). In fact, a biosensor has been developed that allows detection of NTE in whole blood (Makhaeva et al., 2007). 3. Compounds other than pentavalent phosphorus compounds also inhibit NTE, including phosphites and carbamates (Abou-Donia, 1995; Abou-Donia and Lapadula, 1990; Lotti and Moretto, 2005). These, however, do not have a leaving group as do the OP compounds with a pentavalent phosphorus atom, and inhibition of NTE is not irreversible so OPIDN that includes both the clinical and the pathological changes described previously has not been reported. 4. OP-induced NTE inhibition occurs in tissues of animal species that do not show the clinical and pathological manifestations of OPIDN obvious in the U.S. EPA hen model. This includes chicks, rats, mice, and Japanese quail (Ehrich, 1996; Funk et al., 1994a; Padilla and Veronesi, 1985, 1988; Varghese et al., 1995; Veronesi et al., 1991; Wu and Casida, 1996). It appears, for instance, that chickens need to be 55–60 days of age before clinical signs indicative of OPIDN appear. Some pathological evidence of OPIDN, however, can be seen in the spinal cord of chicks that were only 2 weeks old when exposed to DFP, suggesting that age susceptibility in this animal model of OPIDN needs reevaluation (Funk et al., 1994a). OP-induced NTE inhibition occurs in mice, but clinical effects are seen earlier, death is likely, and the pathological picture is not similar to that seen in hens and rats with OPIDN (Wu and Casida, 1996). NTE has been purified, cloned, and localized in neuronal cell bodies (Chang et al., 2008a,b; Glynn, 2006; Glynn et al., 1993, 1994, 1998, 1999; Johnson and Glynn, 2001; Li et al., 2003; Lush et al., 1998; Mackay et al., 1996; Thomas et al., 1993). Its tertiary structure has been

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suggested (Wijeyesakere et al., 2007). Purification, characterization, and physiological effects of OP phosphorylation and aging of NTE have been aided by use of new, very potent probes and sensitive substrates (Borhan et al., 1995; Wu and Casida, 1994, 1995; Yoshida et al., 1994, 1995). Studies using cultured neural cells, yeast, Drosophila (with Swiss cheese protein, NTE’s fruit fly homolog), and heterozygous NTE knockout mice have reported that NTE acts on membrane lipids and has lysophospholipase and phospholipase B activity (Quistad et al., 2003; Read et al., 2007). It deacetylates phospholipids of intracellular membranes (Glynn, 2005, 2006; Zaccheo et al., 2004) and hydrolyzes lysophosphatidylcholine (Casida and Quistad, 2005; Casida et al., 2008; Vose et al., 2007, 2008). Glycerolphosphocholine production is decreased, phosphatidylcholine homeostasis is disrupted, and neuronal and glial cells die when NTE activity is reduced (Muhlig-Versen et al., 2005; Zaccheo et al., 2004). The Drosophila Swiss cheese protein has been reported to interact with a subunit of cAMP-activated protein kinase C, another possible means by which it can contribute to neurodegeneration (Bettencourt da Cruz et al., 2008). Studies with Drosophila mutants showing neurodegeneration have suggested that degenerative diseases other than OPIDN could occur when NTE is depleted (Kretzschmar, 2005), and a report has related NTE status to motor neuron disease in humans (Rainier et al., 2008). Homozygous loss of NTE or its analog, Drosophila Swiss cheese mutant, causes lethality of mice, and heterozygous mutant mice had developmental abnormalities (Moser et al., 2004). Pathological changes have been reported in NTE-deficient mice, even in the absence of OP exposure (Akassoglou et al., 2004), but they do not resemble lesions of OPIDN, and as noted previously, mice are not susceptible to clinical and neuropathological changes that occur with classical OPIDN (Wu and Casida, 1996). Furthermore, although inhibition of NTEassociated phospholipases occurred in nervous tissue from OP-treated adult hens that went on to develop OPIDN, no change in levels of phosphatidylcholine or lysophosphatidylcholine were noted (Hou et al., 2008). Also, although inhibition of the enzymes lysophospholipase and phospholipase B and decrease of glycerophosphocholine occurred in nervous tissue from mice given the NTE inhibitor tri-o-cresyl phosphate (which “aged” the enzyme), no alterations in nervous system levels of phosphocholine and lysophosphatidylcholine occurred (Hou et al., 2009). These studies further suggest that the connection between NTE inhibition and the morphological evidence of OPIDN is not necessarily direct. Because NTE does not appear to be the single target of neuropathy-inducing OP compounds, phosphorylation of other cellular proteins has been suggested to be associated with the development of OPIDN (O’Callaghan, 2003). These include other serine esterases and hydrolases (Casida and Quistad, 2005), signaling proteins and protein kinases, and cytoskeletal proteins such as neurofilaments, tubulin, and actin (Carlson and Ehrich, 2001; Fox et al., 2003; Gupta

Hayes’ Handbook of Pesticide Toxicology

and Abou-Donia, 1994, 1995a,b; Hargreaves et al., 2006). Alteration of neurotrophins has also been suggested (Pope et al., 1995), but this did not occur in chickens exposed to PSP (Pomeroy-Black et al., 2007). As demonstrated in vitro, cytoskeletal changes occur in cells exposed to neuropathy-inducing OP compounds, although the specificity of these changes is unclear. For example, these studies have demonstrated decreased neurite extension in the presence of neuropathy-inducing OP compounds (Hargreaves et al., 2006; Hong et al., 2003; Massicotte et al., 2003), although increased neurite extension has been reported after exposure of neuronal cells to tissue extracts from hens with OPIDN (Pope et al., 1995). Furthermore, decreased neurite extension in cultured neuronal cells has not always been seen following exposure to OP compounds inducing neuropathy, and other studies have reported that cholinesterase-inhibiting OP compounds that do not cause OPIDN also decreased neurite extension (Chang et al., 2005, 2006; Das and Barone, 1999). In vitro and in vivo studies have been conducted in the search for specific cytoskeletal proteins affected by the presence of the neuropathy-inducing OP compounds TOCP, a protoxicant, or DFP, which causes both OPIDN and cholinergic poisoning. Radiolabeled OP compounds were found to bind to proteins other than NTE, with increase in phosphorylated proteins seen early rather than later (Abou-Donia, 1993; 1995; Carrington and Abou-Donia, 1985; Patton et al., 1985). The investigators proceeded to attempt to identify the protein(s) phosphorylated following exposure to a neuropathy-inducing OP, TOCP. Gel electrophoresis indicated that exposure of hens to TOCP enhanced in vitro phosphorylation of tubulin, microtubule-associated protein-2, and neurofilament proteins of 70, 160, and 210 kDa. These are cytoskeletal proteins important in the maintenance of axonal integrity. Because only Ca2calmodulin kinase II activity catalyzes the phosphorylation of these cytoskeletal proteins, effects of neuropathy-inducing OP compounds on this enzyme were also examined (Lapadula et al., 1991, 1992; Suwita et al., 1986a,b), and calmodulin binding was found to increase. The investigators suggested that the hyperphosphorylation of cytoskeletal proteins decreases their ability to be transported down the axon, causing accumulation (Abou-Donia, 1993; Gupta et al., 1997). The preceding studies measured phosphorylation of cytoskeletal proteins in vitro in tissues from animals exposed to neuropathy-inducing OP compounds. To verify if excess phosphorylation of neurofilaments actually occurred in nervous tissue, immunohistochemical techniques were used to determine the status of phosphorylated neurofilaments in affected myelinated nerve fibers of hens exposed to neuropathyinducing OP compounds (Jensen et al., 1992). This study demonstrated an excess accumulation of phosphorylated neuro­ filaments in swollen axons at 21 (distal sciatic nerve) or 7 (spinal cord dorsal columns) days after administration of 750 mg/kg protoxicant TOCP. However, immunohistochemical studies in another laboratory did not demonstrate prominent excessive phosphorylated neurofilament aggregates prior to

Chapter | 69  Organophosphorus-Induced Delayed Neuropathy

fiber degeneration in susceptible axonal populations of hens given a neuropathic dosage of the active NTE inhibitor PSP (Jortner et al., 1999). In vivo studies using 2 mg/kg intramuscular PSP, which initiates OPIDN without causing acetylcholin­ esterase-inhibiting toxicity, demonstrated that expression of -tubulin was altered in spinal cord of hens within 48 h of exposure, days before development of OPIDN (Fox et al., 2003). The PSP-induced effect on tubulin verified studies done with DFP, a potent cholinesterase inhibitor, but provided a much more restricted list of gene expression than seen with the latter (Damodaran et al., 2000, 2001; Gupta et al., 1999). For example, downregulation of neurofilament expression was not seen in nervous tissue from hens given PSP, although it was reported in hens given DFP or TOCP (Gupta et al., 1999; Zhao et al., 2006). Instead, neurofilament degradation has been reported (Song et al., 2009). In contrast to the in vivo studies, studies in neuroblastoma cell cultures did not reveal change in total -tubulin but, rather, demonstrated a transient rise in phosphorylated neurofilaments, the 200-kDa neurofilament protein and the phosphorylated forms of the signaling protein ERK 1/2 (Cho and Tiffany-Castiglioni, 2004; Hargreaves et al., 2006). Another study with neuronal cells in culture demonstrated significant decreases in the cytoskeletal protein F-actin early with micromolar concentrations of three of four neuropathy-inducing OP compounds tested, although exposure to millimolar concentrations of non-neuropathy-inducing paraoxon and parathion for longer periods of time (24 h) also affected concentrations of this cytoskeletal protein (Carlson and Ehrich, 2001). Biochemical changes noted when neuronal cells were exposed to neuropathy-inducing OP compounds included alterations of the membrane transmembrane potential and ATP production (Carlson and Ehrich, 1999; Massicotte et al., 2005). Detrimental effects on ATP were also noted in peripheral nerve of hens given PSP, an effect correlated with decreases in nerve conduction velocity (Massicotte et al., 2001). The loss of ATP and the other alterations discussed previously could be contributors to the electrophysiological alterations and change in axonal transport reported earlier. For example, a number of studies, although not all (Chemnitius et al., 1988; Shell et al., 1988), indicated that administration of neuropathy-inducing OP compounds altered nerve and/or muscle electrophysiological responses in humans (Roberts, 1977; Vasilescu et al., 1984), hens (Anderson et al., 1988; Durham and Ecobichon, 1984; El-Fawal et al., 1988, 1989, 1990b,c; Lidsky et al., 1990; Robertson et al., 1987, 1988), dogs (Schaeppi et al., 1984), and cats (Abou-Donia et al., 1986; Baker et al., 1980; Drakontides and Baker, 1983; Lapadula et al., 1982). Specifically, an increase in threshold excitability was noted in peripheral nerves (sciatic, tibial, and biventer cervicis) of the hen, which is the accepted animal model for OPIDN early (1–4 days) after administration of the OP neurotoxicants PSP and DBDCVP (El-Fawal et al., 1988, 1989, 1990b,c; Robertson et al., 1987, 1988).

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Morphological damage was seen in the biventer cervicis nerve as early as 4 days after administration of PSP, even though clinical manifestations of OPIDN were not evident before postdosing day 10 (El-Fawal et al., 1990c). Alteration of axonal transport following administration of neuropathy-inducing OP compounds has also been investigated. Studies in the cat sciatic nerve indicated that anterograde axonal transport was accelerated 7 days after administration of DFP, an effect the investigators suggested was secondary to the pathologic effect because the change was relatively small (Carrington et al., 1989). More recent studies from this laboratory, however, suggest that sciatic nerve transport of neurofilament protein in the hen is first increased (at 3 days) and then decreased (7 days) after exposure to DFP (Gupta et al., 1997). Another laboratory examined retrograde transport in hens given DBDCVP. Retrograde transport decreased in the ventral spinal cord 3 days after treatment. Maximal effects were noted 7 days after dosing, with effects both in the ventral spinal cord and in the dorsal root ganglia (Moretto et al., 1987). Clinical deficits were not reported before day 10 after dosing.

69.7  Factors influencing the development of organophosphorus-induced delayed neuropathy As noted previously, both inhibition and “aging” of NTE are expected before OPIDN will occur. OP compounds that do not age do not cause OPIDN (Johnson, 1982). In fact, nonaging inhibitors of NTE, given prior to neuropathy-inducing OPs, will prevent OPIDN. These NTE inhibitors include carbamates, phosphinates, and sulfonyl fluorides. These compounds appear to protect the NTE from the OP compound; that these compounds protect this enzyme has been a primary reason for designating NTE as the primary target for initiation of OPIDN (Johnson, 1982; Johnson and Read, 1993). Promotion or potentiation, in which administration of certain NTE inhibitors after neuropathy-inducing OP compounds will initiate or exacerbate clinical manifestations and nervous system lesions of OPIDN, has been reported in many studies (Johnson and Read, 1993; Lotti, 2002; Lotti and Moretto, 1999; Massicotte et al., 1999; Moretto et al., 2005; Pope and Padilla, 1990; Pope et al., 1992, 1995; Randall et al., 1997). The exacerbation of OPIDN appears to be due to a quantitative rather than qualitative difference, as observed in hens given several different OP compounds (e.g., DFP, DBD-CVP, and PSP) and several different NTE inhibitors, with phenyl methanesulfonyl fluoride being used most often (Massicotte et al., 1999; Moretto et al., 1992a; Osman et al., 1996; Peraica et al., 1995; Randall et al., 1997). To date, all promotors of OPIDN are NTE inhibitors, yet most are those that do not lose a side group after attachment to the enzyme (in other words, the promoter–enzyme complex does not have to age)

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(Moretto et al., 2005). During promotion, OPIDN can appear at subclinical doses of a neuropathy-inducing OP compound or in test subjects normally not susceptible to this condition (e.g., chicks and rats) (Harp et al., 1997; Lotti et al., 1993, 1995; Moretto et al., 1992a,b; Pope et al., 1992, 1993, 1995). The only enzyme consistently inhibited by promoters was NTE. However, NTE may not be the single target of OPIDN for promotion (Gardiman et al., 1999; Lotti, 1992, 1997; Lotti and Moretto, 1999, 2005; Lotti et al., 1993; Moretto et al., 2005; Osman et al., 1996; Pope et al., 1993). Factors other than NTE inhibition may be involved in OPIDN and promotion of OPIDN, because a soluble factor released in the spinal cord after exposure to a neuropathic OP compound had dramatic effects on cell growth (Pope et al., 1995). Although a number of events that occur between OP exposure and development of OPIDN have been identified, questions remain about susceptibilities. It is evident that certain species (e.g., the accepted animal model, the hen, as well as humans, cats, ferrets, sheep, dogs, and turkeys) are more susceptible than others (e.g., rats and mice) to clinical manifestations of the toxicity (Johnson, 1982). Certain strains of hens, the animal model of OPIDN, also appear to be differentially affected (Bursian et al., 1989; Dunnington et al., 1989; Ehrich et al., 1986a). Furthermore, young animals appear to be relatively resistant to clinical manifestations of toxicity (Funk et al., 1994a; Johnson, 1982; Lotti and Moretto, 2005; Moretto et al., 1991; Peraica et al., 1993). NTE inhibition, however, can be significant in most susceptible and nonsusceptible species and age groups. In addition, although different and less extensive than that seen in the hen, neuropathological manifestations can be noted in populations once thought not to be susceptible to OPIDN (e.g., rats and young chicks) (Ehrich et al., 1995; Funk et al., 1994a; Padilla and Veronesi, 1985). Comparative studies on the effects of neuropathyinducing OPs in hens and rats have been done, including studies of the progression and regression of lesions (Carboni et al., 1992; Dyer et al., 1991, 1992; Ehrich et al., 1995; ElFawal et al., 1990c; Jortner et al., 1989). These studies were done with hens 18 months old and rats more than 60 days old. Indications were that lesions could repair in both species over time, with repair occurring considerably earlier in the rat than in the hen. The reason for this is that in hens, repair is manifest only in peripheral nerve, where myelinated nerve fiber regeneration is seen over a period of weeks following fully developed OPIDN (Jortner et al., 1989). The repair of single exposure-induced lesions in the rat is thought to be related to return of swollen, vacuolated fasciculus gracilis axons to the normal state – a more rapid process (Carboni et al., 1992). It has been suggested that such capability for repair could at least partially explain species and age differences to induction of OPIDN, including the low susceptibility of chicks and increased susceptibility of older rats (6 months) as measured by clinical evidence of neurological damage following treatment with neuropathyinducing OP compounds (Funk et al., 1994a; Lotti, 1992, 2002; Moretto et al., 1992b; Peraica et al., 1993).

Hayes’ Handbook of Pesticide Toxicology

Although repair may play a role in age and species susceptibilities to clinical and pathological manifestations of OPIDN, this process may be less significant in the treatment of OPIDN. Changes in manifestations of OPIDN have been examined in animal models of this disorder (hen, cat, and rat) using stress situations or therapeutic agents that have been suggested to exaggerate or provide therapeutic advantage, respectively, in people or experimental animals with natural or experimentally induced neurological disorders (Capildeo, 1989; McEwan, 2000; Wilkinson, 1993). Stress has been suggested to contribute to adverse health outcomes in a variety of situations, including exposures to environmental chemicals (Institute of Medicine, 2003). Social stress exaggerated clinical signs in hens exposed to TOTP (Ehrich and Gross, 1983), but using corticosterone administration as a surrogate for stress along with multiple exposures to both a neuropathy-inducing OP (TOTP) and a nonneuropathic OP (chlorpyrifos) did not exaggerate OPIDN in the rat (Ehrich et al., 2004; Jortner et al., 2005). Glucocorticoids, which are elevated during times of stress, are used in the treatment of patients with acute traumatic injuries of the nervous system (Capildeo, 1989; Wilkinson, 1993). The first studies examining the potential of glucocorticoids to ameliorate OPIDN (induced by DFP) were done in cats (Baker and Stanec, 1985; Baker et al., 1982; Drakontides et al., 1982). The depression of repetitive neural discharges and muscle contractile response usually seen in cats with OPIDN did not appear, and morphological damage to motor nerve terminals was much attenuated. Glucocorticoids could also ameliorate OPIDN in hens, an effect that was dependent on dose of both corticoid and OP compound, with relatively low concentrations of glucocorticoids protective, as indicated by clinical, electrophysiological, and morphological endpoints (Ehrich and Gross, 1982; Ehrich et al., 1986b, 1988; Lidsky et al., 1990; Piao et al., 2004). When doses were higher, glucocorticoids (and extreme stress) could exacerbate OPIDN without effect on NTE (Ehrich and Gross, 1986; Ehrich et al., 1985, 1986a,b, 1988) (Figure 69.16). Calcium channel blockers are also used to treat neuro­ logical disorders, especially those related to ischemia (Brailowsky, 1988). The rationale for studies on amelioration of OPIDN is based on the general role of calcium in neuronal degradation (Schlaepfer and Hasler, 1979; Schlaepfer and Zimmerman, 1984). Axonal degeneration, which is a feature of OPIDN, has been suggested to result from an increase in calcium-dependent proteinase (CANP or calpain) activity (Schlaepfer and Zimmerman, 1984). Calpain activity increased in brain, sciatic nerve, and muscle of hens treated with TOTP or PSP, with activity significantly increased in sciatic nerve as early as 2 days after treatment with PSP (El-Fawal et al., 1990a). Total nerve calcium was also increased, with this effect noted 4 days after PSP treatment. Increases in calpain activity were blocked by administration of 4 daily doses of the calcium channel blocker nifedipine when initiated 1 day before PSP treatment. Calcium channel blockers

Chapter | 69  Organophosphorus-Induced Delayed Neuropathy

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4.0 TOTP and 300 ppm corticosterone TOTP and 200 ppm corticosterone

Clinical score Mean ± SD N = 8-12

3.0

TOTP alone

TOTP and 50 ppm corticosterone

2.0

TOTP and 30 ppm corticosterone 1.0

10

12 15 18 Days after TOTP

21

Figure 69.16  Effect of 30–300 ppm corticosterone on clinical signs of delayed neuropathy induced by po administration of TOTP (360 mg/kg) to chickens. Results are presented as mean  SD, n  8–12. A score of 0  no clinical signs, 1  mild ataxia, 2  moderate ataxia, 3  severe ataxia, and 4  paralysis. Chickens given corn oil or corticosterone without TOTP had scores of 0. Reproduced with permission from Ehrich, M., Jortner, B. S., and Gross, W. B. (1986). Dose-related beneficial and adverse effects of dietary corticosterone on organophosphorus-induced delayed neuropathy in chickens. Toxicol. Appl. Pharmacol. 83, 250-260. Copyright © 1986 by Academic Press, San Diego.

were demonstrated to ameliorate PSP- and TOTP-induced OPIDN, as indicated by clinical, electrophysiological, and morphological endpoints (El-Fawal and Ehrich, 1993; El-Fawal et al., 1989, 1990a,b). Antibodies to proteins indicative of neurodegeneration were also less in serum of hens with OPIDN that had been treated with calcium channel blockers (El-Fawal and McCain, 2008). Clinical signs developed later in hens treated with the calcium channel blockers verapamil and nifedipine. In addition, excitability thresholds of nerve–muscle preparations from hens given PSP and calcium channel blockers approached levels in preparations from control animals, and the pathological effects of PSP on myelinated fibers of the biventer cervicis nerve were markedly attenuated (Figures 69.17 and 69.18). Calcium channel blockers have also been demonstrated to decrease toxicity of DFP, another neuropathy-inducing OP compound (Dretchen et al., 1986). Other reports, however, noted less OPIDN in hens treated with calcium-containing products (Muzardo et al., 2008; Piao et al., 2003). The amelioration of OPIDN by calcium channel blockers may be related to their effects on calpain (El-Fawal and Ehrich, 1993; El-Fawal et al., 1990a) or to their action against differential vascular effects induced by neuropathic and nonneuropathic OP compounds (McCain et al., 1993, 1995). Calcium channel blockers did not affect NTE. Neuropathy-inducing PSP increased peripheral vascular resistance, response to vasoactive agents, and circulating levels of norepinephrine and epinephrine. The calcium channel blocker verapamil attenuated all of these responses.

The effects of PSP on the cardiovascular system did not occur in hens exposed to paraoxon, an OP compound that does not cause OPIDN, suggesting that OP effects on the cardiovascular system may contribute to development of OPIDN (McCain et al., 1993).

69.8  Testing for organophosphorusinduced delayed neuropathy Registration of OP compounds for pesticide use under FIFRA recommends that they be tested in hens 8–14 months old without designation of breed or strain. Since the 1990s, this testing has included NTE and acetylcholinesterase determinations, clinical observations, and neuro­ pathology following single- and multiple-dosing procedures (U.S. EPA, 1996). In the initial testing procedure, brain and spinal cord samples are collected from a subset of the dosed hens within 48 h of administration of a single dose of the test OP insecticide, and NTE and acetylcholinesterase activities are determined. The remaining hens are observed over the next 3 weeks, with in situ perfusion–fixation and removal of brain, spinal cord, and peripheral nerves for histopathological examination at that time. Multiple-dose testing (28 days) may also be necessary. With these tests, the relative sensitivity of NTE to inhibition compared to acetylcholinesterase inhibition identifies those OP compounds capable of causing OPIDN even before clinical signs and morphological changes appear. Although testing requirements have been unchanged during the past decade,

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12 11 10 9 8

5 *

* *

3 2

*

*

*

* PSP VP

*

P

7 Strength (v)

Clinical score

4

6 5 4

NP

VP C NP

1

7

A

10

13

15

18

21

Days after administration

24

27

30

40

B

60

80 100

Duration (µ sec)

Figure 69.17  (A) Development of clinical deficits and partial recovery after administration of phenyl saligenin phosphate (PSP), nifedipine plus PSP (NP), or verapamil plus PSP (VP). Verapamil, 7 mg/kg/day im, was given for 4 days beginning 1 day before PSP, 2.5 mg/kg im. Nifedipine, 10 mg/ kg/day, was given for 5 days beginning 1 day before PSP. Results are presented as mean  SE, n  5–10. Differences between the group of hens given only PSP and hens given nifedipine or verapamil plus PSP are denoted by asterisks (ANOVA with Newman–Keuls test for multiple comparisons, p  0.05). 1  altered gait, 2  difficulty in walking and standing, 3  severe ataxia, 4  leg paralysis, and 5  paralysis with both leg and wing involvement. Hens not given PSP had scores of 0. (B) Log–log plot for inflection region (40–500 s) of strength–duration curves from biventer cervices nerve muscle preparation days 15 and 16 after treatment of hens with PSP (2.5 mg/kg im). The dosing regimen is as described for Figure 69.17A. C, control; NP, nifedipine plus PSP; P, PSP; and VP, verapamil plus PSP. Reproduced with permission from El-Fawal, H. A., Jortner, B. S., and Ehrich, M. (1990). Modification of phenyl saligenin phosphate-induced delayed effects by calcium channel blockers: In vivo and in vitro electrophysiological assessment. Neurotoxicology 11, 573-592. Copyright © 1990 by Intox Press, Little Rock, AR.

Figure 69.18  Cross sections of distal levels of the biventer cervicis nerve from chickens dosed with 2.5 mg/kg im of phenyl saligenin phosphate 15 days earlier. The nerve in A is from a hen that only received the toxicant and shows extensive loss of myelinated fibers. The nerve in B was from a hen that had received the toxicant plus the calcium channel blocker verapamil at 7 mg/kg/day for 4 days beginning 1 day prior to the phenyl saligenin phosphate administration. Examination of this nerve (B) shows that the verapamil dosing was protective to the myelinated nerve fibers, many of which have a normal morphological appearance (arrows). Toluidine blue–safranin stain; scale bar  100 m.

Chapter | 69  Organophosphorus-Induced Delayed Neuropathy

Conclusion OPIDN is a generally progressive, irreversible disorder that causes clinical manifestations appearing days to weeks after humans and certain species of animals are exposed to OP compounds that can essentially irreversibly inhibit most of the available NTE (neurotoxic esterase). The severity of OPIDN, as indicated by clinical and pathological manifestations, depends on species and age of test animals and the extent of NTE inhibition. Chickens have proven to be the most sensitive test species. OPIDN is manifest clinically by ataxia and weakness progressing to paralysis, associated with bilateral degeneration of distal and terminal regions of long myelinated nerve fibers. The neuropathy can be prevented by pretreatment with NTE inhibitors; however, these

AChE and NTE activities

mouse AChE mouse NTE

% of control

100

human AChE human NTE

80 60 40 20 0 10–9

10–8

A

10–7

10–6

10–5

10–4

10–3

Paraoxon (M)

AChE and NTE activities

human AChE mouse AChE

100 % of control

recent concern about isomeric forms of the test compounds has been raised (Battershill et al., 2004). Suggestions have been made that NTE measurements in cultured cells could be used to predict potential for OPIDN without the need to run this test in animals (Barber et al., 1999a,b, 2001; Ehrich and Veronesi, 1999; Ehrich et al., 1997; Funk et al., 1994b,c; Hong et al., 2003; Knoth-Anderson and Abou-Donia, 1993; Knoth-Anderson et al., 1992; Nostrandt and Ehrich, 1992, 1993; Qian et al., 2007; Sogorb et al., 1997). Investigations indicated that NTE activity could be found in both primary cultures of avian and bovine origin (e.g., dorsal root ganglia neurons, hen embryo brain spheroids, and chromaffin cells) (Massicotte et al., 2003, 2005; Quesada et al., 2007; Sales et al., 2004) and continuous cell lines of human and rodent origin (e.g., SH-SY5Y, PC-12, and NB41A3) (Ehrich and Veronesi, 1999; Ehrich et al., 1997). A thorough concentration–response study with 11 active esterase inhibitors, including 7 that cause OPIDN and 4 that do not, indicated that either a human cell line or a murine cell line was capable of identifying the neuropathy-inducing OP compounds based on the relative sensitivity of NTE to inhibition compared to acetylcholinesterase (Ehrich et al., 1997) (Figure 69.19). Concentrations of OP compounds needed to inhibit NTE and acetylcholinesterase were far below those cytotoxic to the cultures. A similar result was noted in another study in which very sensitive NTE inhibitors were examined in cell lines of rodent origin (Li and Casida, 1997). Although cell cultures did not have sufficient oxidative capability to convert protoxicant phosphorothioates to active enzyme inhibitors (Ehrich and Veronesi, 1999; Ehrich et al., 1997), later studies indicated that this can be overcome by preincubation of OP protoxicants with a bromine solution or a microsomal preparation (Barber et al., 1999a,b). Cell cultures could also be used to determine NTE inhibition following 28 days of exposure (Barber and Ehrich, 2001). The results of recent studies enhance the possibility that OP compounds may one day be screened for potential to induce OPIDN by using an in vitro system, but one must always use caution when extrapolating in vitro findings to in vivo situations (Ehrich and Dorman, 2004).

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human NTE mouse NTE

80 60 40 20 0 10–8

B

10–7

10–5

PSP (M)

Figure 69.19  Concentration–response curves for inhibition of acetylcholinesterase (AChE) and neuropathy target esterase (NTE) in neuroblastoma cells of human and murine origin by organophosphorus compounds. Cells were incubated with OP compounds for 1 h before assay. Paraoxon causes acute cholinergic crisis (AChE inhibition) rather than organophosphorusinduced delayed neuropathy (OPIDN); PSP causes OPIDN. Point-to-point composite curves are provided to aid visualization (Prism; GraphPad, San Diego). Each curve represents at least three different assays that included at least three concentrations of OP compounds that provided values between 10 and 90% of values in vehicle-treated cells. Reproduced with permission from Ehrich, M., Correll, L., and Veronesi, B. (1997). Acetylcholinesterase and neuro­ pathy target esterase inhibitions in neuroblastoma cells to distinguish organophosphorus compounds causing acute and delayed neurotoxicity. Fundam. Appl. Toxicol. 38, 55–63. Copyright © 1997 by Academic Press, San Diego.

same compounds promote OPIDN when given after a neuro­ pathy-inducing OP compound. The temporal order of precise mechanisms of OPIDN has not been determined, but changes in cellular calcium homeostasis and/or cytoskeletal proteins may be involved because these changes appear in vitro and in vivo after exposure to neuropathy-inducing OP compounds.

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Shell, L., Jortner, B. S., and Ehrich, M. (1988). Assessment of organophosphorus-induced delayed neuropathy in chickens using needle electromyography. J. Toxicol. Environ. Health 25, 21–33. Sogorb, M. A., Bas, S., Gutierrez, L. M., Vilanova, E., and Viniegra, S. (1997). Bovine chromaffin cells as an in vitro model for the study of non-cholinergic toxic effects of organophosphorus compounds. Arch. Toxicol. Suppl. 19, 347–355. Somkuti, S. G., Tilson, H. A., Brown, H. R., Campbell, G. A., Lapadula, D. M., and Abou-Donia, M. B. (1988). Lack of delayed neurotoxic effect after tri-o-cresyl phosphate treatment in male Fischer 344 rats: Biochemical, neurobehavioral, and neuropathological studies. Fundam. Appl. Toxicol. 10, 199–205. Song, F., Yan, Y., Zhao, X., Zhang, C., and Xie, K. (2009). Neurofilaments degradation as an early molecular event in tri-ortho-cresyl phosphate (TOCP) induced delayed neuropathy. Toxicology 258, 94–100. Stumpf, A. M., Tanaka, D. J., Aulerich, R. J., and Bursian, S. J. (1989). Delayed neurotoxic effects of tri-o-tolyl phosphate in the European ferret. J. Toxicol. Environ. Health 26, 61–73. Suwita, E., Lapadula, D. M., and Abou-Donia, M. B. (1986a). Calcium and calmodulin stimulated in vitro phosphorylation of rooster brain tubulin and MAP-2 following a single oral dose of tri-o-cresyl phosphate. Brain Res. 374, 199–203. Suwita, E., Lapadula, D. M., and Abou-Donia, M. B. (1986b). Calcium and calmodulin-enhanced in vitro phosphorylation of hen brain coldstable microtubules and spinal cord neurofilament triplet proteins after a single oral dose of tri-o-cresyl phosphate. Proc. Natl. Acad. Sci. USA 83, 6174–6178. Tanaka, D. J., and Bursian, S. J. (1989). Degeneration patterns in the chicken central nervous system induced by ingestion of the organophosphorus delayed neurotoxin tri-ortho-tolyl phosphate: A silver impregnation study. Brain Res. 484, 240–256. Tanaka, D., Bursian, S. J., and Lehning, E. (1990). Selective and terminal degeneration in the chicken brainstem and cerebellum following exposure to bi(1-methylethyl)phosphorofluoridate (DFP). Brain Res. 519, 200–208. Tanaka, D., Bursian, S. J., and Lehning, E. J. (1992). Neuropathological effects of triphenyl phosphite on the central nervous system of the hen. Fundam. Appl. Toxicol. 18, 72–78. Tanaka, D. J., Bursian, S. J., Lehning, E. J., and Auterich, R. J. (1991). Delayed neurotoxic effects of bis(1-methylethyl) phosphorofluoridate (DFP) in the European ferret: A possible mammalian model for organophosphorus-induced delayed neurotoxicity. Neurotoxicology 12, 209–224. Thomas, T. C., Szekacs, A., Hammock, B. D., Wilson, B. W., and McNamee, MG. (1993). Affinity chromatography of neuropathy target esterase. Chem. Biol. Interact. 87, 347–360. U.S. Environmental Protection Agency (EPA) (1996). “Health Effects Test Guidelines OPPTS 870.6100. Delayed Neurotoxicity of Organophosphorus Substances Following Acute and 28-Day Exposure.” Available at http://www.epa.gov/opptsfrs/publications/OPPTS_Harmonized/870_Health_Effects_Test_Guidelines/ Drafts/870-6100.pdf  . Varghese, R. G., Bursian, S. J., Tobias, C., and Tanaka, D. Jr. (1995). Organophosphorus-induced delayed neurotoxicity: A comparative study of the effects of tri-ortho-tolyl phosphate and triphenyl phosphite on the central nervous system of the Japanese quail. Neurotoxicology 16, 45–54.

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Vasilescu, C., Alexianu, M., and Dan, A. (1984). Delayed neuropathy after organophosphorus insecticide (Diperterex) poisoning: A clinical, electrophysiological and nerve biopsy study. J. Neurol. Neurosurg. Neuropsychol. 47, 543–548. Veronesi, B. (1984). Effect of metabolic inhibition with piperonyl butoxide on rodent sensitivity to tri-ortho-cresyl phosphate. Exp. Neurol. 85, 651–660. Veronesi, B., and Dvergsten, C. (1987). Triphenyl phosphite neuropathy differs from organophosphorus-induced delayed neuropathy in rats. Neuropathol. Appl. Neurobiol. 13, 193–208. Veronesi, B., Padilla, S., and Lyerly, D. (1986). The correlation between neurotoxic esterase inhibition and mipafox-induced damage in rats. Neurotoxicology 7, 207–216. Veronesi, B., Padilla, S., Blackmon, K., and Pope, C. (1991). Murine susceptibility to organophosphorus-induced delayed neuropathy (OPIDN). Toxicol. Appl. Pharmacol. 107, 311–324. Vose, S. C., Holland, N. T., Eskenazi, B., and Casida, J. E. (2007). Lysophosphatidylcholine hydrolases of human erythrocytes, lymphocytes, and brain: Sensitive targets of conserved specificity for organophosphorus delayed neurotoxicants. Toxicol. Appl. Pharmacol. 224, 98–104. Vose, S. C., Fujioka, K., Gulevich, A. G., Lin, A. Y., Holland, N. T., and Casida, J. E. (2008). Cellular function of neuropathy target esterase in lysophosphatidylcholine action. Toxicol. Appl. Pharmacol. 232, 376–383. Wijeyesakere, S. J., Richardson, R. J., and Stuckey, J. A. (2007). Modeling the tertiary structure of the patatin domain of neuropathy target esterase. Protein J. 26, 165–172. Wilkinson, I. M. S. (1993). “Essential Neurology”. Blackwell, Oxford. Winrow, C. J., Hemming, M. L., Allen, D. M., Quistad, G. B., Casida, J. E., and Barlow, C. (2003). Loss of neuropathy target esterase in mice links organophosphate exposure to hyperactivity. Nature Genet. 33, 477–485. Wu, S. Y., and Casida, J. E. (1994). Neuropathy target esterase inhibitors: Enantiomeric separation and stereospecificity of 2-substituted-4H1,3,2-benzodioxaphosphorin 2 oxides. Chem. Res. Toxicol. 7, 77–81. Wu, S. Y., and Casida, J. E. (1995). Ethyl octylphosphonofluoridate and analogs: Optimized inhibitors of neuropathy target esterase. Chem. Res. Toxicol. 8, 1070–1075. Wu, S. Y., and Casida, J. E. (1996). Subacute neurotoxicity induced in mice by potent organophosphorus neuropathy target esterase inhibitors. Toxicol. Appl. Pharmacol. 139, 195–202. Yoshida, M., Wu, S. Y., and Casida, J. E. (1994). Reactivity and stereospecificity of neuropathy target esterase and a-chymotrypsin with 2-substituted-4H-1,3,2-benzodioxaphosphorin 2 oxides. Toxicol. Lett. 74, 164–176. Yoshida, M., Tomizawa, M., Wu, S. Y., Quinstad, G. B., and Casida, J. E. (1995). Neuropathy target esterase of hen brain: Active site reactions with 2-[octyl-3H]octyl-4H-,3,2-benzodioxaphosphorin 2 oxide and 2-octyl-4H-, 3,2-[aryl-3H]benzodioxaphosphorin 2 oxide. J. Neurochem. 64, 1680–1687. Zaccheo, O., Dinsdale, D., Meacock, P. A., and Glynn, P. (2004). Neuropathy target esterase and its yeast homologue degrade phosphatidylcholine to glycerophosphocholine in living cells. J. Biol. Chem. 279, 24024–24033. Zhao, X. L., Zhang, T. L., Zhang, C. L., Han, X. Y., Yu, S. F., Li, S. X., Cui, N., and Xie, K. Q. (2006). Expression changes of neurofilament subunits in the central nervous system of hens treated with tri-orthocresyl phosphate (TOCP). Toxicology 223, 127–135.

Chapter 70

Chlorpyrifos Emanuela Testai, Franca M. Buratti and Emma Di Consiglio Istituto Superiore di Sanità, Rome, Italy

70.1  Introduction and regulatory aspects Chlorpyrifos (CPF) is an organophosphorothionate (OPT) insecticide with nonsystemic anticholinesterase activity with contact, stomach, and respiratory action. It has been widely used in agriculture, horticulture, viticulture, and forestry on a wide range of crops, in residential and nonresidential applications to control cockroaches, fleas, ticks on cattle, and pests in animal houses. The U.S. Environmental Protection Agency (EPA) estimates that in the United States more than 20 million pounds of CPF are used yearly (U.S. EPA, 2002) of which approximately 11 million pounds are applied in nonagricultural settings. The extensive use is due to its effective, costcompetitive broad spectrum of activity when compared with alternative products, and for this reason CPF has often been chosen as a replacement for persistent organochlorinated compounds. Chlorpyrifos volatilization is indeed a significant dissipative process in the environment and leaching is not relevant; therefore, there is negligible risk of following crops or groundwater contamination. Nowadays CPF is registered in over 100 countries, including both industrialized and developing nations; it is the most studied of the OPTs, with more than 2000 studies and reports published that evaluate the impact of the active ingredient on the environment and human health, especially in children. Indeed, in the past years an excess of children’s dietary intake caused by food-to-surface-to-mouth or surface-to-hand-to-food activities has been shown (Adgate et al., 2001; Fenske et al., 2002a); consequently, in 2002 the EPA confirmed the decision to eliminate all homeowner use and to reduce nonagricultural uses to less than 3%, limiting them to mosquito control in public health (U.S. EPA, 2002). Furthermore, the EPA indicated measures for farmers’ use to mitigate the occupational and the ecological risks (e.g., adopting a 24-h waiting period before entering fields where CPF has been applied) (U.S. EPA, 2002). In Canada the Pest Management Regulatory Agency has implemented measures

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to mitigate risk for people and the environment similarly to the EPA (PMRA, 2003). The Australian Pesticides and Veterinary Medicines Authority permits the use of products containing CPF with limitations, such as CPF limit concentrations (50 g/l) in liquid formulations sold for domestic use and indoor applications (NRA, 2005). In the European Union (EU), CPF is included in Annex I to Directive 91/414/EEC, allowing products containing CPF as the active ingredient to be reregistered by the regulatory authorities of individual EU member states (EU, 2005). Both agricultural and residential uses of large quantities of CPF are in place in many developing countries of the Pacific and Latin America areas and in Africa. Hundreds of thousands of deaths per year have been reported in these areas due to acute poisoning episodes stemming from excessive use in agriculture, poor use of adequate individual protection devices (e.g., gloves and protective clothes), and intentional ingestion (Eddleston et al., 2007; Eddleston and Phillips, 2004; Yasmashita et al., 1997).

70.2  Physicochemical properties The compound is a colorless to white crystalline solid, has a mercaptan-like odor, and is only slightly soluble in water but is soluble in most organic solvents. The most important physicochemical properties are reported in Table 70.1. Chorpyrifos degrades slowly in soil under both aerobic and anaerobic conditions. It has a half-life of 60–120 days, degrading to nontoxic 3,5,6-trichloro-2-pyridinol (TCP), which is subsequently degraded to organochlorine compounds and carbon dioxide. Spray drift to surface water may occur, but CPF dissipates in microbially active natural water systems with a half-life of less than 1 week and tends to migrate to sediments; therefore, the risk for aquatic organisms is limited by the short-lived exposure (WHO, 2007). The environmentally acceptable concentration adopted by WHO is 1 g/l, even in the event of multiple applications.

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Table 70.1  Physicochemical Properties Chemical name

O,O-diethyl-O-(3,5,6-trichloro2-pyridyl)phosphorothioate

Molecular weight

350.6

Empirical and structural formula

C9H11Cl3NO3PS

CAS registry number

2921-88-2

State

Crystalline solid

Color

White to tan

Odor

Mild mercaptan

Melting point

41.5–42.5°C

Boiling point

.300°C

Vapor pressure

3.35 mPa at 25°C

Density

1.51 g/ml at 21°C

Solubility, mean

Acetone 400 g/l at 20°C Dichloromethane 400 g/l at 20°C Ethyl acetate 400 g/l at 20°C Methanol 250 g/100 ml at 20°C Toluene 400 g/l at 20°C n-Hexane 400 g/l at 20°C Water 1.05 ppm (w/v at 25°C)

Partition coefficient (n-octanol and water)

Kow  50,000

Stability

(a) In sterile buffered water pH 4.7 at 25°C: half-life of 63 days pH 6.9 at 25°C: half-life of 35 days pH 8.1 at 25°C: half-life of 23 days (b) In organic solvents No signs of degradation in xylenerange aromatic solvents used in CPF formulations

70.3  Toxicokinetic properties 70.3.1  Absorption Chorpyrifos can be absorbed through the ­ gastrointestinal mucosa, lung epithelium, and skin. After oral exposure CPF is quantitatively absorbed by the gastrointestinal tract and enters the bloodstream. When 50 mg/kg body weight (bw) CPF is administered to rats in corn oil, .80% of the dose was absorbed, resulting in cholinesterase (ChE) inhib­ ition (Timchalk et al., 2002b); by using an intestinal perfusion model, 99% of CPF was absorbed by rat small intestine (Cook and Shenoy, 2003). The absorption is rapid; indeed, blood Cmax of the three metabolites diethylphosphate (DEP), diethyl thiophosphate (DETP), and TCP was detected 3 h after CPF administration (Timchalk et al., 2002b). When rats were treated orally with the three metabolites (140 mol/kg bw),

they also were quantitatively absorbed within 3 h (Timchalk et al., 2007). After oral doses of 1 and 10 mg/kg bw, CPF blood levels were independent of age in weanling rats (5, 12, and 17 days old) (Timchalk et al., 2006). The absorption is similar in humans and rat (Fagerholm et al., 1996): the rat and human jejunum effective intestinal permeability highly correlate, and both can be used to predict in vivo oral absorption in humans. Human volunteers dosed with a single oral intake of 0.5 and 1 mg CPF/kg bw showed 70% and 93% absorption, respectively, measured as TCP excreted in the urine (Griffin et al., 1999; Nolan et al., 1984). Absorption could be even greater when taking into consideration other elimination pathways such as bile and feces and the possible protein binding. Differences may also occur due to the formulation and/or physical form of CPF used, as suggested by a more recent study showing that only 20–35% of the administered dose (0.5, 1, or 2 mg/kg bw) was recovered as TCP in the urine of 12 human volunteers (Timchalk et al., 2002b). Chorpyrifos is considered to be absorbed through the respiratory tract in humans and animals, breathing sprays or dust (Geer et al., 2004). Although the actual extent of absorption and rate has not been quantified, its occurrence is well demonstrated by the acute effects evidenced in experimental animals after acute inhalation exposure (Berteau and Deen, 1978; Corley et al., 2009; Dow, 1983). In rats exposed to up to 0.295 mg CPF/m3, no inhibition of red blood cell (RBC) cholinesterases was evidenced (Corley et al., 2009). In humans, average airborne CPF concentrations ranging from 10 to 1100 g/m3 resulted in 19–32% butyrylcholinesterase (BuChE) inhibition in 175 workers involved in the production of CPF (Brenner et al., 1989). The indoor range of CPF air concentration of 0.001–0.1 g/m3 does not inhibit plasma BuChE or RBC acetylcholinesterase (AChE) (Byrne et al., 1998). The percutaneous absorption of CPF has been evidenced in animal studies by induced toxic effect after dermal expos­ ure. A value of 3% can be obtained by the ratio of the oral and dermal lowest observed adverse effect levels (LOAELs) (0.3 mg/kg/day from the rat developmental neurotoxicity study and 10 mg/kg/day from the 21-day rat dermal study) (U.S. EPA, 2002). In humans a similar range between 1 and 3% of the CPF applied dose for dermal absorption was determined, based on TCP recovery in the urine (Griffin et al., 1999; Nolan et al., 1984). Since the proportion of the administered dose metabolized to TCP is unknown, these estimates are considered minimum values (i.e., absorption could be higher), and higher values up to 10% have been suggested (Krieger, 1995; Mage, 2006). In addition, not all of the absorbed dose is eliminated as urinary metabolites, as evidenced in human volunteers dermally dosed with CPF (5 and 15 mg): 4.3% was recovered in the urine as TCP without dose dependence and the nonabsorbed amount was 42–73%. The CPF clearance was not completed within 120 h, suggesting that CPF (or TCP) could be retained in the skin at the application site and/or by other tissues (Meuling et al., 2005). In vitro 52.15 mM CPF dissolved in water or

Chapter | 70  Chlorpyrifos

in ­ ethanol had different dermal penetration rates, indicating a role for the vehicle; after 24 h a significant amount of the applied dose remained bound to the skin, acting as a reservoir for releasing CPF over a longer period (Griffin et al., 2000).

70.3.2  Distribution No study on the distribution of CPF and its metabolites in humans is available. However, animal studies give some indications. Based on octanol: water partition coefficients and lipid content of tissues, Timchalk calculated CPF distribution between various tissues and blood: fat (435:1)  brain (33:1)  liver (22:1)  kidney (10:1) (Timchalk et al., 2002b). Although chlorpyrifos oxon (CPFO) is less lipophilic than the parent, the same trend was found (Timchalk et al., 2002b). Consequently, CPF and CPFO seems to accumulate in the adipose tissue, although in this calculation plasma protein binding and the influence of biotransformation are not considered. Indeed, the rates of biotransformation and metabolite excretion largely inhibit bioaccumulation and consequently mitigate bioconcentration along the food chain.

Figure 70.1  Chlorpyrifos biotransformation pathway.

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In pregnant animals CPF can pass the placenta and reach the fetus: 14C-labeled CPF given to pregnant rats is distributed (as CPF and its metabolites) to all maternal and fetal tissues and plasma (Abdel-Rahman et al., 2002); radio­ activity traces were detected also in placenta and amniotic fluid (Akhtar et al., 2006). In humans the presence of CPF and metabolites in postpartum meconium was evidenced, suggesting absorption across the umbilical cord and diffusion across the placenta surface (Whyatt and Barr, 2001). The elimination of CPF and TCP is slow, the rates probably depending on redistribution from lipid stores (AbdelRahman et al., 2002).

70.3.3  Metabolism The major pathways of CPF metabolism are shown in Figure 70.1. They include: Oxidative desulfuration of the PS moiety to PO, catalyzed by cytochrome P450 (CYP), resulting in the toxic intermediate CPFO (bioactivation)

l

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Dearylation catalyzed by CYP, resulting in 3,5,6-trichloro2-pyridinol (TCP) and diethyl thiophosphate (DETP) (detoxification) l Hydrolysis by A-esterases (paraoxonases-PON1) of the phosphate ester bonds of CPFO to form TCP and diethylphosphate (DEP) (detoxification) l Hydrolysis by B-esterases as aliesterases (carboxylesterase-CE) and cholinesterase (BuChE), acting as molecular scavengers by binding stoichiometrically to CPFO (detoxification) l Conjugation of CPFO by glutathione-S-tranferases with reduced glutathione (GSH) (detoxification) l Conjugation of TCP by glucuronil-transferases and sulfotransferases to form the corresponding glucuronide and sulfate conjugates (detoxification)

l

CPF is a weak acetylcholinesterase (AChE) inhibitor per se but it can be desulfurated by several isoenzymes of the CYP family to form the phosphate triester CPFO (McBain et al., 1971; Sultatos et al., 1984a,b; Sultatos, 1994), which is a powerful inhibitor of brain and serum AChE (Forsyth and Chambers, 1989). CYPs can also catalyze CPF dearylation/ dealkylation, which is considered a detoxication reaction, giving rise to TCP and DETP (Figure 70.1) (Sultatos et al., 1982; Sultatos, 1994). During the desulfuration reaction, the formation of activated sulfur atoms, able to bind ­irreversibly to the active CYP, causes enzyme loss and reduction of the corresponding monooxygenase activity (Halpert et al., 1980). This can influence the metabolism of both endogenous compounds as steroid hormones (Usmani et al., 2003, 2006) and other chemicals to which it is possible to be coexposed (Hodgson and Rose, 2007), such as other pesticides (fipronil and carbaryl) (Joo et al., 2007; Tang et al., 2002), diesel fuel component (nonane) (Joo et al., 2007), and drugs (imipramine) (Di Consiglio et al., 2005). The major site of CPF metabolism is the liver; however, extrahepatic metabolism has been reported in other tissue as brain (Chambers and Chambers, 1989; Pond et al., 1995) and intestine (Poet et al., 2003). Intestinal cells also express on their membrane P-glycoproteins, which have CPFO as a substrate: by transporting it into the lumen they could play an additional role in CPF detoxification (Lanning et al., 1996). Comparing CPF metabolism between rat brain and liver, hepatic microsomes were .1000-fold more active (Chambers and Chambers, 1989; Pond et al., 1995). CYP metabolic rate in enterocytes exceeds that in liver microsomes on a nmolP450 basis, whereas on a mg-protein basis the hepatic microsomes were 61-fold and two-fold more active for TCP and CPFO, respectively (Poet et al., 2003). Human, rat, and mouse liver microsomes form the detoxification metabolite TCP more readily than CPFO; overall, human hepatic microsomes were less active than rodents in the formation of both metabolites (Tang et al., 2001). As shown by in vitro studies with both human liver microsomes (HLM) and c-DNA expressed enzymes, the CYPs involved in the dearylation/ dealkylation

Hayes’ Handbook of Pesticide Toxicology

are CYP1A2, 2B6, 3A4, 3A5, 3A7, 1A1, 2C19, 2C9, and 2D6 (Buratti et al., 2003, 2006; Mutch and Williams, 2006; Sams et al., 2004; Tang et al., 2001). CYP2B6 seems to have the highest catalytic efficiency in producing CPFO, while CYP2C19 is mainly involved in CPF dearylation (Mutch and Williams, 2006; Sams et al., 2004; Tang et al., 2001); other CYPs form variable levels of two metabolites (Buratti et al., 2003, 2006; Mutch and Williams, 2006; Sams et al., 2004; Tang et al., 2001), depending on the CPF dose. Indeed, by considering the average CYP content in the human liver, the isoenzymes that mainly activate CPF to CPFO at CPF concentrations 10 M (considered as representative of actual exposure conditions to low environmentally relevant levels) are CYP1A2 and CYP2B6 (Buratti et al., 2003), whereas at higher CPF concentrations (.100 M), typical of accidental acute intoxication or poisoning episodes, CYP3A4 and CYP2B6 are more efficient (Buratti et al., 2003; Tang et al., 2001). Recently, the catalytic activity of recombinant human fetal liver CYP3A7 has been investigated: only TCP formation was evidenced at 5 M CPF, whereas at high substrate levels (50 M) CYP3A7 can form both metabolites, with TCP formation always higher than CPFO. The efficiency of the fetal isoform in vitro toward CPF was similar to CYP3A4 and CYP3A5, the major adult isoforms (Buratti et al., 2006). Considering that the CYP3A family (3A4 and 3A5) accounted for 82% of intestinal isoforms (Paine et al., 2006), together with the presence of the 2C family (2C9 and 2C19) (Obach et al., 2001; Zhang et al., 1999), the relevance of human intestine as a site for extrahepatic CPF metabolism is evident. Some of the active CYPs, such as 2B6, 1A2, and 2C19, are polymorphic enzymes; therefore, genetic polymorphism can be one of the causes of differing individual susceptibilities to CPF-induced effects (Ingelman-Sundberg, 2002). Furthermore, CYP1A2 can be easily induced by common habits such as smoking or by consumption of caffeine­containing beverages and well-done meat (aromatic amines). Chlorpyrifos itself can induce the human CYPs 1A1, 1A2, 2B6, and 3A4 (Rose and Hodgson, 2005). Consequently, individuals endowed with higher levels of CYP2B6 and CYP1A2 activities could produce CPFO more efficiently than others when exposed to low levels of CPF, possibly experiencing higher levels of ChE inhibition. On the other hand, CYP2C19 allelic variants, mainly involved in detoxication, showed significant decrease in dearylation efficiency with respect to the wild-type forms, possibly affecting in vivo CPF toxicity (Tang et al., 2001). Together with the direct dearylation/dealkylation of CPF catalyzed by different CYPs, the hydrolysis of CPFO to inactive compounds by plasma oxonases (PON1) (Costa et al., 2005a, Li et al., 1995; Pond et al., 1995) represents the main detoxication pathway (Figure 70.1). PON1 is active on CPFO and is not able to hydrolyze directly the parent CPF (Costa et al., 2005a). PON1 is a polymorphic enzyme: alleles, indicated as Q192 and R192, show different catalytic

Chapter | 70  Chlorpyrifos

efficiencies toward specific substrates. R192 is more efficient with CPFO than Q192 allozyme, whereas the two alloforms show the same level of activity toward the diazinon-oxon; therefore, while R192 presence provides protection toward CPF toxicity, no difference is evidenced after diazinon expos­ ure (Costa, 2006). The development of a transgenic mouse model expressing the different polymorphic human PON1 (Furlong et al., 1998) and the use of PON1 knockout mice (Furlong, 2007) have confirmed the effect of PON1 polymorphism. Not only the allele, but also the level of expression are important to evidence individual susceptibility; indeed, in a given population plasma PON1 activity can vary up to 40-fold and differences in PON1 protein levels up to 13-fold are also present within a single genotype (Costa et al., 2005a; Costa, 2006). Furthermore, there is a characteristic ethnic distribution of PON1 activity: in Finns, Europeans, and Indians the frequency of low PON activity is high (Diepgen and Geldmacher-von Mallinckrodt, 1986), an excess of subjects with high activity has been reported in Nigerian populations, and an unimodal distribution due to prevalence of individuals with intermediate activity (QR) has been described among Mexicans (Rojas-Garcia et al., 2005). In addition pharmacological or nutritional status can increase PON1 activity to a limited extent (Costa et al., 2005b). Human hepatic R192 PON1 showed Km values of 0.25 mM for CPFO (Costa et al., 2003). In the conditions of low level actual human exposure, CYP1A2 is active (Km  0.95  103 mM) (Buratti et al., 2002), but the levels of CPFO are too low when compared to PON1 Km values; therefore, the role of this enzyme in detoxication could be limited. Timchalk et al. utilized a Monte Carlo analysis and a physiologically based pharmacokinetic and pharmacodynamic (PBPK/PD) model to investigate the impact of genetic polymorphism of human PON1 on the theoretical concentration of CPFO in the brain (Timchalk et al., 2002a,b). At low CPF doses (5 g/kg) the model is insensitive to changes in CPF-oxonase. However, with an increasing dose (500 g/kg) the model suggests a dose-dependent nonlinear increase in CPFO concentration in the brain, associated with PON1 activity. After repeated treatment at low doses (5 g/kg), the brain CPFO concentrations were comparable to the one obtained after a single exposure, suggesting that at low envir­onmentally relevant exposures PON1 activity can be compensated by other esterase catalyzing detoxification pathways (Timchalk et al., 2002a), such as B-esterases (BuChE and CE). Carboxylesterases bind stoichiometrically to CPFO and very efficiently act as alternative targets to AChE, decreasing its concentration in blood (Chambers and Carr, 1993; Kutz et al., 1992; Pond et al., 1995; Talcott et al., 1979b). They are a multigene family widely distributed in mammalian tissue, with the highest levels expressed in the liver (in the endoplasmic reticulum and the cytosol), the gastro­ intestinal tract, and the brain (Satoh et al., 2002; Satoh and Hosokawa, 1998). In humans, the serum CE is apparently

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lacking (Li et al., 2005), as opposed to the rat, where the CE activity in plasma is quite high, exerting a more evident protective effect toward CPF-induced effects. Gender and age differences have been observed in rat CE activity: female rats have higher CE levels (Chanda et al., 1997) as well as adults compared with weanling animals, with consequent differences in the response to CPF exposure (Karanth and Pope, 2000). Two major human CE forms have been characterized: hCE1 and hCE2 in the liver (Satoh et al., 2002) and one (hCE2) in the intestine (Imai et al., 2006), but which one is involved in CPF detoxification has not been clarified yet. CE activity in human liver has a moderate level of interindividual variability (about two- to four-fold) (Buratti and Testai, 2005; Ross and Crow, 2007). No significant variation between populations of different ethnic origin has been evidenced (Hosokawa et al., 1995; Talcott et al., 1979a). Their levels of expression are constant over the lifetime (Pope et al., 2005), but possible variation in CE activity can be due to single-nucleotide polymorphisms, which have been recently described in the human liver (Hosokawa et al., 1990; Marsh et al., 2004), and by induction phenomena, mediated by a number of drugs and a variety of compounds, including the product of the oxidative OPT desulfuration, acting by decreasing the enzymatic activity (Satoh and Hosokawa, 1998). Since a significant number of other pesticides and pharmacologically active drugs are metabolized by CE, alteration in their activity to massive binding to CPFO may be the cause of significant toxicological interactions. The predominant cholinesterase in human serum is BuChE (.99%), which is consequently the primary defense against CPF toxicity: its inhibition has been used as biomarker of exposure to CPF together with the urinary TCP detection and RBC AChE activity. When data on these three biomarkers were collected in CPF-exposed workers over a year-long study (1999–2000), it was found that plasma BuChE activity was negatively related to urinary TCP concentrations (Garabrant et al., 2008). Absorbed CPF doses of 5 and 50 g/kg/day (well below the general population exposure levels) were identified as no-effect levels for inhibition of plasma BuChE and RBC AChE, respectively (Garabrant et al., 2008), although the high intraindividual variability in cholinesterase activities measurements in nonexposed workers (used as controls) showed that monitoring programs can have a substantial rate of biased results (Garabrant et al., 2008). The conjugation between GSH and CPF or CPFO mediated by glutathione-S-transferase (GST) has been evidenced by the presence of glutathione-derived metabolites in the urine of an acute CPF-poisoned patient (Bicker et al., 2005b). In addition, glutathione-derived metabolites have been detected in in vitro studies with human hepatocytes and pooled S9 liver (Choi et al., 2006). Recently, GSTs have been demonstrated to catalyze CPF and CPFO O-deethylation to five GSH conjugates, but it is not clear yet which human GST is involved (Fujioka and Casida, 2007).

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Glucuronyl-transferases and sulfotransferases actively conjugate TCP with glucuronic acid and sulfate. These compounds are much more hydrophilic than CPF and can be excreted in the bile and in urine (Bicker et al., 2005b; Choi et al., 2006). Due to the age-related sensitivity to CPF-induced effects and the concern about neurodevelopmental tox­icity related to CPF exposure during pregnancy/early postnatal periods, the biotransformation in different age groups has been extensively investigated. The overall analysis of the available data on fetus, on juvenile rats, and to a more limited extent, on human newborns and young children led the EPA to conclude that the age-related sensitivity to CPF effects can be attributed, at least in part, to toxicokinetic differences (U.S. EPA, 2008a).

70.3.4  Excretion The major CPF elimination pathway is urinary excretion, with TCP as the most abundant product, together with DEP, DETP, GSH conjugates, sulfates, and glucuronides (Bicker et al., 2005a; Nolan et al., 1984). Unmetabolized CPF is not present in the urine. The half-life for elimination of CPF from the various organs in rats is comparable (10–16 h), except for elimination from fat, which was estimated to be 62 h (Smith et al., 1967). The elimination half-life in humans has been estimated to be 27 h, with the maximum rate of TCP excretion occurring 24–48 h following dermal expos­ ure (Nolan et al., 1984). CPF may be also eliminated in the bile/feces as GSH and glucuronide conjugates (Hirom et al., 1972), which may be hydrolyzed in the intestinal lumen and reabsorbed entering the enterohepatic recirculation. Another possible elimination pathway is through the breast milk, which, being relatively rich in fats, can accumulate CPF and CPFO. Low doses of CPF have been detected in rat milk (0.3 mg/kg/day) during gestational studies (Mattsson et al., 2000). In humans, measurable levels of CPF in breast milk from nursing mothers in India were detected, calculating that newborns consuming 500 ml milk/day could be exposed to 2–60 g CPF/kg (Sanghi et al., 2003).

70.4  Exposure Occupational exposure to CPF can occur during mixing, spray application, cleaning up of equipment, or disposal of containers, while the general population is exposed through spray drift, residues present in treated food and drinking water, or indirect exposure from the environment, although its environmental persistence is quite limited. The main pathways of exposure are inhalation of vapors and aerosols, dermal absorption, or ingestion of residues in the diet. Accidental acute exposure to high levels of CPF is associated with the lack of individual protection devices used by agricultural workers or to not following good agricultural

Hayes’ Handbook of Pesticide Toxicology

practices. However, the most common pattern of exposure to CPF is subchronic occupational exposure to moderate CPF levels and long-term low levels of exposure for the general population due to environmental or dietary exposure. A Total Diet Study by the Center of Food Safety and Applied Nutrition of the Food and Drug Administration (FDA) considered the typical diet consumption, in the years 1993–2003, of adults (70 kg body weight), toddlers (age 2 years, 20 kg body weight), and infants (ages 6–11 months, 10 kg body weight). Approximately 6000 food samples were analyzed to detect pesticide residues. Chlorpyrifos was detected in 18% of samples, and a possible CPF dietary intake of 0.005, 0.014 and 0.009 g/kg/day for adults, toddlers, and infants, respectively, was derived (CFSAN, 2001). Recently, many studies have been focused on farmworker families and agricultural area bystanders, who can be exposed by reentry into CPF-treated areas or by handling treated animals (the socalled “take-home” exposure), making inhalation of vapors/ aerosols and dermal absorption potential exposure pathways also for bystanders. These routes of exposure should be limited for the general population, considering that CPF residential use has been phased out in many countries (e.g., in the United States since 2002). However, the studies showed that the oral route seems to be the most relevant since no difference in the urinary TCP levels in farm and “nonfarm” children were reported (Alexander et al., 2006; Curwin et al., 2007a,b; Fenske et al., 2002a,b). The urinary concentration of the primary CPF metabolites in children were up to two times higher than levels observed in comparable studies with adults (Gurunathan et al., 1998), likely due to an excess of dietary intake (Adgate et al., 2001; Fenske et al., 2000, 2002a) and the typical pattern of exposure associated with food-tosurface-to-mouth or surface-to-hand-to-food habits. A recent review on CPF (Eaton et al., 2008) summarized the results from a number of CPF exposure assessment studies, most of which deal with the children cohort studies organized by the EPA to evaluate exposure to CPF and other pesticides in different areas of the United States. From the analysis of the exposure data, most of which was based on urinary TCP, with samples being collected before the U.S. residential use phased out, the authors reported a conservative estimate of CPF current daily exposure for the adult general population of 0.004–0.006 g/kg/day. These data, in line with data obtained by the dietary study by the FDA, confirm that the oral route is the major source of CPF expos­ure, which is below the reference values. The estimate is, however, conservative, since the direct absorption of TCP, being directly present in the environment as a degradation product, likely leads to overestimatation of CPF exposure using TCP urinary excretion as a biomarker. Abiotic hydrolysis, photolysis, and plant metabolism can convert OPT residues to dialkylphosphates (DAPs) not only in the environmental compartment but also on treated fruits and vegetables. For this reason, in control analysis the definition of CPF residue includes the parent compound, TCP,

Chapter | 70  Chlorpyrifos

and the conjugates, all expressed as CPF equivalents (EU, 2005). The extent of these conversions was evaluated by analyzing 153 product samples. A total of 91 out of 153 samples (60%) contained more DAP residues than parent OPTs, supporting the consideration that biomonitoring studies measuring metabolites excreted in urine may be highly affected by dietary ingestion of DAPs (Zhang et al., 2008), leading to overestimations of 10- to 20-fold (Morgan et al., 2005; Wilson et al., 2003). Therefore, CPF dosing in blood is sometimes prefereed (Barr et al., 2002), considering the prolonged excretion rate (24–48 h urine collection after single exposure), which could limit the application of TCP urinary excretion on a routine basis (Maroni et al., 2000). Environmental exposure to CPF has been estimated at about 1.4 g/kg/day in the U.S. adult population (Schurdut et al., 1998). More recently the estimate of the concentration of CPF (or its metabolites) in different conditions has been possible through PBPK/PD models, quantitatively describing the relationships among biological processes following CPF exposure. By modeling data on TCP urinary concentration, the magnitude and timing of CPF exposure doses in children were back-estimated: an average background of 0.15 g CPF/kg/day was predicted in children who experienced exposure events to an estimated dose as low as 1.61 g/kg (Rigas et al., 2001). A PBPK/ PD model has been developed for CPF in rats and humans: the obtained simulations were compared with experimental data of published available data on target tissue dosimetry and dynamic response, following a single oral dose of 0.5– 100 mg/kg for rats and 0.5–2 mg/kg for humans, improving the extrapolation of data on rodents to human (Timchalk et al., 2002b). Other PBPK models were developed to identify biological reference values (BRVs) of urinary biomarkers obtained in biomonitoring programs (generally DAPs) to assess health risk of workers (Bouchard et al., 2005) and children (Valcke and Bouchard, 2009). In the case of children, BRVs for methyl-phosphate and ethylphosphate derivatives of 52 nmol/kg bw and 32 nmol/kg bw for oral or dermal exposure, respectively, were obtained. The altered ratio of CPF levels between plasma and tissues in late pregnancy was also evidenced by models, showing that the plasma partitioning of CPF increased with the higher lipid levels in plasma, with consequently lower levels of CPF in other tissue compartments. Modifying lipid-tissue partition coefficients in PBPK/PD models for rats and humans resulted in plasma CPF levels that were proportional to the changes in plasma lipids during gestation (Lowe et al., 2009). This could be relevant in relation to the in utero exposure of the conceptus. Indeed, indoor and/or outdoor exposure of pregnant women to OPTs (mainly to CPF) has been associated with different birth outcomes (Berkowitz et al., 2004; Eskenazi et al., 2004; Whyatt et al., 2004). In humans, the feasibility of prenatal exposure has been demonstrated by the detection of CPF in umbilical cord blood (Whyatt et al., 2004) and of OPT

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metabolites in human postpartum meconium, the intestinal content of the fetus, very likely due to absorption across the umbilical cord, diffusion across the placenta surface, and/or swallowing by the fetus of amniotic fluid (Whyatt and Barr, 2001). The levels in meconium (0.80–3.20 g/g DEP and 2.0–5.6 g/g DETP) can be considered as a longterm dosimeter for prenatal exposure, reflecting exposures from the second trimester to birth, taking into account the extended half-life of OPT metabolites in the medium (Whyatt and Barr, 2001). This can explain the striking difference with levels measured in TCP in the urine, markers of short-term exposure only. Indeed, pesticide exposure is generally episodic, being characterized by peaks during (or immediately after) application and decreasing levels afterward, due to relatively rapid clearance and low potential of bioaccumulation. More recent studies failed to find CPF in meconium, although the exposure of the mother was detected by CPF in hair (Ostrea et al., 2009), very likely due to lower levels of exposure following the restriction in the use of CPF-containing products.

70.5  Toxicological profile 70.5.1  Mechanism of Action The mechanism of toxicity caused by organophosphorus pesti­cide is the inhibition of AChE, which catalyzes the hydrolysis of the neurotransmitter acetylcholine (ACh) (Klaassen, 2001), as the result of a chemical reaction between the phosphoryl (PO) moiety and a serine hydroxyl group in the active site of AChE, producing an inactive phosporylated enzyme. Since CPF is an OPT (PS), to exert its action it has to be converted, via oxidative desulphuration (Gallo, 1991; McBain et al., 1971; Sultatos et al., 1984a, b; Sultatos, 1994) to CPFO, the active metabolite, which is much more potent than CPF itself as an AChE inhibitor (Huff et al., 1994; Ma and Chambers, 1994). Other enzymes and proteins also have serine amino acids in their active sites or allosterically situated near the active site; therefore, CPF may affect their functions, suggesting the possibility for effects other than AChE inhibition (Barone et al., 2000; Casida and Quistad, 2004; Pope, 1999). The nervous system is the primary target of CPFO acute toxic action. Indeed, AChE phosphorylation results in the accumulation of Ach in the synapses and neuromuscular junctions, leading to overstimulation of cholinergic (i.e., muscarinic and nicotinic) receptors. The excess of ACh causes clinical signs and a spectrum of effects on the CNS, on the autonomic nervous system, and at neuromuscular junctions, depending on the type of affected receptor (Taylor, 1996; Watanabe, 1989). At muscarinic receptor sites, ACh accumulation causes prolonged response, both excitatory (bronchoconstriction) and inhibitory (vasodilatation); at nicotinic receptors, the excitatory effect results

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in involuntary contraction of the muscle fibers whereas the inhibitory effect leads to muscular paralysis. Although AChE spontaneous reactivation is a complex and slow process, AChE inhibition cannot be considered irreversible. However, without an appropriate treatment with antidotes, CPF action could induce long-lasting effects. The most frequently used treatment is with 2-pralidoxime methiodide (2-PAM), binding the phosphorylated AChE, with release of the unbound enzyme. The phosphorylated AChE may also be transformed into a state in which no spontaneous reactivation occurs and oximes are no longer able to reactivate it. This specific mechanism, called aging, involves dealkylation of the phosphoryl groups attached to the enzyme, producing irreversible inhib­ ition. The toxic effects persist until a sufficient amount of new AChE is synthesized to eliminate the excess of the neurotransmitter. AChE is also present in plasma and in red blood cells (RBCs). Inhibition of plasma ChE is not itself considered an adverse effect (U.S. EPA, 2000b); on the contrary, RBC ChE inhibition is considered a critical effect, useful as a marker of potential adverse effects on the nervous system. Indeed, RBC ChE has a close structural affinity to brain ChE, and its inhibition usually precedes the inhibition of brain ChE (U.S. UEPA, 2000a, 2002; van Gemert et al., 2001). Moreover, in rats it has been demonstrated that the measurement of RBC enzyme inhibition is more sensitive than in target tissues, that is, brain, heart, or voluntary muscle: at 10 mg/kg CPF, 90, 56, and 41% inhibition was measured for RBC, heart, and brain ChE, respectively (Nordstrand et al., 1997). In animals 26% acute inhibition of brain AChE is associated with neurobehavioral changes (Tilson et al., 1993). A statistically significant depression (starting from at least 50%) of brain ChE activity is usually associated with cholinergic signs (Bignami et al., 1975; Cochran et al.,

2009), and at least in mice the relationship between brain ChE inhibition and behavior may be considered a continuum in which low CPF doses produce subtle neurobehavioral changes that become evident at higher doses (Cometa et al., 2007). In field workers a 50% inhibition of RBC AChE or 60% inhibition of plasma AChE is needed to induce poisoning signs (Lotti, 1995). Besides AChE inhibition there is also a morphogenic role for this enzyme that may direct the architecture of the central and peripheral nervous systems to cause permanent damage (Bigbee et al., 1999; Brimijoin and Koenigsberger, 1999). Chlorpyrifos inhibits other esterases (or pseudocholinesterase), such as BuChE, which are present in plasma and in the nervous system of some species (Lotti, 1995). Inhibition of BuChE is not linked to any specific physiological effects, does not lead to any toxicological consequences, but exerts a protective action, acting as an alternative target to AChE (Eaton et al., 2008). In a human volunteer study no clinical signs of toxicity were shown also in the presence of 85% BuChE inhibiton (Nolan et al., 1984).

70.5.2  Acute and Subchronic Toxicity There is a significant database of studies on the acute toxicity of CPF. A summary of the acute toxicological profile of CPF is reported in Table 70.2. CPF shows moderate acute toxicity by the oral route: according to the oral LD50 values mice are more sensitive than rats, guinea pigs, and rabbits (Table 70.2) and female are more sensitive than male rats. CPF is moderately toxic following acute dermal and inhalation exposures; it is considered a mild irritant, although it does not meet EU criteria for classification as a skin and eye irritant in rabbit, and it is not a sensitizer to the skin of guinea pigs (Table 70.2).

Table 70.2  Toxicological Profile of Chlorpyrifos, Based on Acute Toxicity, Irritation, and Sensitization Parameter

Species

Results

Reference

Oral LD50 (mg/kg)

Rat Mouse Guinea pig Rabbit

95–270 62.5 500 1000

Gallo (1991); Kidd and James (1991)

Dermal LD50 (mg/kg)

Rat Rabbit

2000 2000

Dow Chemical Co. (1986); Gallo (1991); Kidd and James (1991)

Inhalation LC50 (4 h) (mg/L)

Rat

1.0 (whole body) 5.22 (nose only)

EU (2005) WHO (2007)

Skin irritation

Rabbit

Non irritant

EU (2005)

Eye irritation

Rabbit

Non irritant

EU (2005)

Skin sensitization

Guinea pig

Non sensitizer

Gallo (1991)

Chapter | 70  Chlorpyrifos

Following CPF exposure, independently of the route of administration, clinical signs were typical of cholinesterase inhibition, leading to the typical “cholinergic syndrome.” When the amount of ACh increases in the parasympatic nervous system, toxicity is generally associated with muscarinic receptor stimulation on the effector organs (i.e., salivary glands, heart, eye, respiratory system, gastrointestinal tract, and blood) and signs of toxicity include salivation, lacrimation, dyspnoea, flaccid paralysis, vomiting, piloerection, exophtalmia, and diarrhea. Signs of toxicity associated with stimulation and subsequent desensitization of nicotinic receptors, located in both the autonomic and somatic systems, include tachycardia, hyper­tension, muscle fasciculation, tremors, and flaccid paralysis (Watanabe, 1989). ACh accumulation in the CNS, where both muscarinic and nicotinic receptors are located, results in combined effects including ataxia, lethargy, mental confusion, weakness, convulsions, respiratory failure, coma, and death (Ecobichon, 1996). At high doses CPF can also induce other systemic toxicity, such as body weight loss, decreased food consumption, liver, kidney, and adrenal pathology (U.S. UEPA, 2000a). Neonates and young rats have been reported to be more susceptible to acute high-level exposures of CPF than adult rats (Moser and Padilla, 1998; Pope and Chakraborti, 1992), with up to 10-fold lower NOELs in neonates (based on ChE inhibition) (Zheng et al., 2000). Several hypotheses have been proposed to explain the age-related ChE differences in sensitivity to CPF, including variations in the rate of recovery of ChE (Mortensen et al., 1998) and limited expression of some detoxification pathways (i.e., oxonase, CE) during early development (Li et al., 1996; Mortensen et al., 1996; Pope et al., 2005). Due to concern for increased sensitivity in young animals, the FQPA (Food Quality Protection Act) Safety Factor Committee recommended in 1999 that an FQPA additional safety factor (10) was needed for the protection of infants and children to exposure resulting from the use of chlorpyrifos (Cleveland et al., 2001). Mice seem to be more susceptible than rats to acute CPF treatment (oral LD50 4.5-fold lower) (Cometa et al., 2007); a good correlation was observed between brain ChE inhib­ ition and classical cholinergic signs of toxicity, and an acute NOEL  12.5 mg/kg was identified (Cometa et al., 2007) that is about eightfold lower than the ones identified for rats in acute oral studies (Clegg and van Gemert, 1999). The species-related differences were less evident after repeated short-term exposure; 25 mg/kg/day could be considered as the LOAEL value and 6.25 mg/kg/day the threshold dose for the appearance of functional signs of CPF-induced anticholinergic activity, both in the rat and in the mouse species (Clegg and van Gemert, 1999; Cometa et al., 2007; Zheng et al., 2000). An overall oral subchronic NOAEL  1 mg/kg bw/day has been derived from 90-day studies with rats, mice, and dogs (EU, 2005; WHO, 2007).

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The differences in acute toxicity could be primarily due to known variations in A-esterases among species (Aldridge, 1993), whereas in repeated treatment the phenomenon of tolerance may play some role in making the CPF-induced effect quantitatively similar in the mouse and rat. Moreover, after repeated treatment significant recovery (up to 70%) of brain ChE activities has been reported in mice within 7 days after treatment (Cometa et al., 2007), whereas there was only a minimal recovery in subcutaneously treated rats (Pope et al., 1991). This is likely due to differences in toxicokinetics between the two routes of administration, the CPF absorption after subcutaneous injection being a prolonged phenomenon, providing a depot of CPF, with production of longer-lasting ChE inhibition as a possible initiating event in muscarinic receptor downregulation (Nordstrand et al., 1997; Pope et al., 1992). In mice the tolerance could not simply be attributed to an alteration in the P450-mediated CPF hepatic metabolism, although some metabolic compensation mech­ anism, involving CE and glutathione-S-transferases, should be in place (Cometa et al., 2007). CPF is moderately toxic to humans (U.S. EPA, 1989), for whom the oral and dermal exposures are the most relevant routes of entry, although plasma ChE activity has been shown to be inhibited also after CPF inhalation (U. S. Public Health Service, 1995). Available information has been derived from a number of reports reviewing the potential risks associated with accidental/unintentional exposures to CPF in domestic or urban environments during the past years. Several human case study reports have been published describing serious poisonings due to inappropriate use by operators or bystanders and to accidental exposures to CPF of adult or children via the oral or the inhalation route (Aiuto et al., 1993; Blondell, 2000). Moreover, it has been recognized that OPT self-poisoning is a global health problem, on the order of hundreds of thousands of deaths per year, mainly in developing countries (Eddleston and Phillips, 2004; Langley and Sumner, 2002). The acute effects in humans do not correlate with animal toxicity. Indeed, although the ranking of OPT acute toxicity (based on oral rat LD50 values) is CPF . fenthion  dimethoate . diazinon . malathion (Guo et al., 2006), recent clinical studies showed that CPF is less toxic in humans than dimethoate and fenthion (Eddleston et al., 2005). Acute poisoning affects the CNS, the cardiovascular system, and the respiratory system. Symptoms in humans occur when ChE activity has been reduced by about 50% and are correlated with effects induced in the CNS or in peripheral organs and tissues and with the typical signs of “cholinergic toxicity” opposite to one another, that is, bradycardia (activation of parasympathic nervous system) and tachycardia (via the sympathetic system). Clinical signs include numbness, tingling sensations, coordination problems, headache, dizziness, tremor, nausea, abdominal cramps, sweating, lacrimation, salivation, blurred vision, difficulty breathing or respiratory depression, slow heartbeat, muscle fasciculation,

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and in some cases the onset of polyneuropathy (Blondell, 2000; Eaton et al., 2008; Sherman, 1995). Very high doses may result in unconsciousness, convulsions, and death, due to either respiratory or cardiovascular failure. The interval between exposure and the onset of symptoms varies from a few minutes to 1–2 h. In 20–50% of acute poisoning cases, CPF causes delayed effects (“intermediate syndrome”), 1–4 weeks after exposure, with or without immediate symptoms (Gallo, 1991). Acute CPF exposure has been reported to produce signs of neurological toxicity weeks or months after the initial symptoms have resolved (Sherman, 1995). Improvement may occur over months or years, although in some cases residual impairment remains (Gallo, 1991). The measurement of RBC ChE could be diagnostic to estimate the severity of the acute CPF-induced intoxication as well as to estimate the need for an antidote therapy (i.e., atropine or oxime products). Oxime therapy has been demonstrated to be efficient against CPF intoxication when compared with other OPTs; indeed, a lower case fatality (8%) induced by CPF has been demonstrated following the standard oxime protocol treatment with respect to the ones induced by other OPTs (e.g., dimethoate, fenthion) (Eddleston et al., 2005). A few human studies with controlled acute and shortterm (28 days) CPF exposure (range of dose tested: 0.014– 5 mg/kg/day) are available (Coulston et al., 1972; Nolan et al., 1984). Based on the available data, humans appear to be as sensitive as rodents to short-term repeated CPF tox­icity, expressed as ChE inhibition (Coulston et al., 1972). In a more recent study, fasting human males and females were administered a single oral CPF dose (0.5, 1.0, or 2.0 mg/kg bw): no clinical signs or symptoms have been reported and a significant RBC cholinesterase inhibition has been evidenced in 1/12 subjects exposed at 2.0 mg/kg, with no significant gender difference (Kisciki et al., 1999), indicating a NOEL of 1 mg/ kg bw. On this basis the 1999 Joint Meeting on Pesticides Residues (JMPR) allocated an acute reference dose (ARfD) of 0.1 mg/kg bw, incorporating a safety factor of 10 to take into account possible intraspecies differences (WHO, 2007). The same result has been obtained by deriving the ARfD value from NOAEL of the 90-day rat study, applying a safety factor of 100 (EU, 2005). The value adopted by the EPA is lower, taking into account an additional protection factor for children and females 13–50 years old (U.S. UEPA, 2000a). Some OPTs are able to induce a noncholinergic delayed neuropathy, named OP-induced delayed neuropathy (OPIDN), in experimental animals and in humans following acute poisoning or after subchronic exposure (Richardson, 1995). The syndrome is characterized by degeneration of sensory and motor axons in distal regions of peripheral nerves, due to the inhibition of the nonspecific esterase known as neuropathy target esterase (NTE), a protein unrelated to AChE and located in the nervous tissue and lymphocytes. In order to induce OPIDN, NTE has to be phosphorylated and then an alkyl side-chain removed completely in the process of “aging” (Richardson, 1995; Sultatos, 1994). Studies carried

Hayes’ Handbook of Pesticide Toxicology

out with hens, the model species for the evaluation of OPIDN, showed that clinical neuropathy was associated with a 50–90% (depending on the OPT) inhibition of NTE (Lotti et al., 1991). The great majority of studies on different animal species have shown that repeated CPF exposure did not induce OPIDN (Richardson, 1995). Only in a few studies did hens experience signs of prolonged locomotor ataxia at CPF doses inducing frank cholinergic toxicity (Anon, 1996; Capodicasa et al., 1991). Signs of OPIDN have been reported in humans, some weeks after massive intentional exposure (Lotti et al., 1986), accidental poisonings (Aiuto et al., 1993), or occupational subchronic exposure (Kaplan et al., 1993). However, in all these positive cases severe cholinergic signs likely associated with extremely high CPF exposure were present, requiring extensive and aggressive atropine/oxime antidotal therapy and artificial ventilation. The weight of evidence led both the EPA (U.S. UEPA, 2000a) and EU (EU, 2005) to conclude that CPF is not responsible for any delayed neuropathy.

70.5.3  Chronic Toxic Effects The effects of concern related to chronic CPF exposure are associated with the inhibition of ChE activity. Data on CPF chronic toxicity are available in rats, mice, and dogs, and independently of the animal species, the inhibition of plasma ChE was the most sensitive parameter. However, its toxicological meaning has been questioned, whereas inhibition of RBC and brain ChE is considered the toxicologically relevant end points, responsible for the onset of clinical signs. A summary of NOELs and critical effects is given in Table 70.3. No substantial differences among species were shown, although beagle dogs appear to be slightly more sensitive than rodents with respect to RBC cholinesterase inhib­ ition, following chronic 2-year CPF dietary exposure. Beside ChE inhibition, as critical effects, at the highest dose tested increased liver weight was also reported in dogs (McCollister et al., 1974); decreased body weight gain and increased adrenal glands weight, characterized microscopically by an exacerbated fatty vacuolation of the zona fasciculate, was observed in rats (McCollister et al., 1974); and reductions in body weight and food consumption, eye opacity, and centrilobular fatty vacuolation of hepatocites were reported in mice (Gur et al., 1991). In contrast to acute toxicity results, following repeated CPF exposure no significant age-related difference in susceptibility to CPF has been evidenced in rats after oral treatment (Zheng et al., 2000). Although over the years most of the attention has been focused on acute poisoning, long-term exposure to low CPF levels involves not only workers but also the vast majority of the population, and the understanding of CPF-induced long-term effects is of great relevance for the protection of human health. The adverse effects of OPTs, including CPF, following long-term low-level exposure have been the subject of debate and the subject is still ­ controversial

Chapter | 70  Chlorpyrifos

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Table 70.3  Summary of Critical Effects of Repeat Dosing with Chlorpyrifos in Laboratory Animals Type of study

Species

Critical effects

LOEL (mg/kg/day)

NOEL (mg/kg/day)

Reference

Oral: 80 weeks

Mouse

Plasma ChE RBC and brain ChE

0.7 6.1

— 0.7

Gur et al. (1991)

Oral:104 weeks

Rat

RBC and plasma ChE Brain ChE

1 3.0

0.1 1.0

McCollister et al. (1974)

Oral: 104 weeks

Rat

Plasma ChE RBC and brain ChE

1.0 10

0.1 1

Young and Grandjean (1988)

Oral: 104 weeks

Rat

RBC and plasma ChE Brain ChE

1.0 10

0.1 1

Yano et al. (2000)

Oral: 104 weeks

Dog

Plasma ChE RBC ChE Brain ChE

0.03 0.1 3.0

0.01 0.03 1.0

McCollister et al. (1974)

(Ray, 2000). Indeed, some studies have failed to show any long-lasting effects in the exposed population (Ray, 1998), whereas other papers reported subtle, mainly cognitive, long-term effects due to chronic exposure to low OPT levels (Beach et al., 1996), and covalent binding to brain proteins, sensitive to low levels of OPTs (Richards et al., 1999). Considering CPF’s mode of action, it is reasonable to hypothesize that slow reversible inhibition of ChE, undetectable after a single dose, could result in the accumulation of subtle effects not relevant in acute or short-term exposure. Indeed, a measurable change in plasma and RBC ChE levels has been repeatedly described in workers exposed to CPF for long periods (American Conference of Governmental Industrial Hygienists, 1986; Dyer et al., 2001; McBain et al., 1971; Yeary et al., 1993). Effects reported in workers chronically exposed to CPF included impaired memory and concentration, disorientation, severe depressions, irritability, confusion, headache, speech difficulties, delayed reaction times, insomnia, and sometimes delayed symptoms. An influenza-like condition with headache, nausea, weakness, and loss of appetite has also been reported (U.S. Public Health Service, 1995). Although symptomatic pesticide poisoning episodes can be followed by persistent cognitive deficits, including lower performance in verbal and visual attention, visual memory, and motor slowing (Roldan-Tapia et al., 2006; Savage et al., 1988; Steenland et al., 1994), only a few cases of patients with memory impairment and intellectual functioning decline have been reported after subchronic asymptomatic exposure to low levels of CPF (McBain et al., 1971) without acute cholinergic episodes (Jamal et al., 2002). Cognitive and motor deficits in adult rats were evident after acute CPF exposure (Bushnell et al., 1993, 2001; Sanchez-Amate et al., 2001) but were less pronounced or absent following repeated exposure to

CPF (Bushnell et al., 1994; Mattsson et al., 1996; Maurissen et al., 2000b; Samsam et al., 2005; Sanchez-Santed et al., 2004). Overall, these findings support the hypothesis that repeated exposure to CPF induces a kind of “tolerance” to inhibition of ChE, likely due to a downregulation of ACh receptors with reduced synthesis of the neurotransmitter ending recovery of the normal function (Bushnell et al., 1993; Richardson, 1995; Sultatos, 1994). Considering the available data, the JMPR adopted an acceptable daily intake (ADI) of 0.01 mg/kg bw/day, derived from the overall NOAEL of 1 mg/kg bw/day based on the inhibition of brain AChE activity in rats, mice, and dogs, using a 100-fold safety factor, or the NOAEL of 0.1 mg/kg bw/day for RBC AChE inhibition in humans, using a 10fold safety factor (WHO, 2007). A similar approach has been followed by EU regulators (EU, 2005), whereas a lower value has been adopted by the EPA, aimed at protecting susceptible groups of the population, that is, children and females 13–50 years old (U.S. UEPA, 2000a).

70.5.4  Genotoxicity and Carcinogeneticity A large number of genotoxicity studies with CPF on bacteria and mammalian cells in vitro and in vivo elicit negative responses (Gollapudi et al., 1995). The few positive studies showing mutagenicity (Patnaik and Tripathy, 1992; Water et al., 1980), sister-chromatid exchanges (Amer and Aly, 1992; Sobti et al., 1982), chromosomal loss (Woodruff et al., 1983), mitotic abnormalities in rats (Roy et al., 1998), and DNA strand breaks in mice (Rahman et al., 2002) were interpreted as the result of an indirect mechanism, related to CPF-induced reactive oxygen species formation (Bagchi et al., 1995), or alterations in endogenous antioxidants

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(Bebe and Panemangalore, 2003). However, the weight of evidence clearly indicates that CPF does not exhibit any genotoxic potential (EU, 2005; U.S. UEPA, 2000a). No evidence of carcinogenicity for CPF has been shown in experimental studies carried out both with rats and with mice. Indeed, there was no treatment-related increase in the incidence of tumors when Fischer F344 rats or CD-1 mice were fed up to 10 and 15 mg/kg/day, respectively, for 104– 105 weeks (Warner et al., 1980; Yano et al., 2000; Young and Grandjean, 1988). Among many of the human epidemiological studies available, a few indications of a potential association between CPF exposure and cancer risk have been reported. Increased risk of lung cancer among applicators exposed to CPF in the United States has been reported, also after adjusting for different lifestyles (i.e., smoking, other occupational exposures, diseases, and dietary habits) (Lee et al., 2004). The possible proposed mechanism is linked to the potential of CPF to induce oxidative stress (Bagchi et al., 1995; Bebe and Panemangalore, 2003) or decrease the activity of glutathione S-transferase (da Silva et al., 2004). Significant correlations have been also found between CPF exposure and rectal cancer (Lee et al., 2007) and increased risk of non-Hodgkin lymphoma among male farmers in the United States (Waddell et al., 2001). However, these findings are based on only a small number of cases and the source and quantitation of exposure are often uncertain, and therefore the possible association between CPF exposure and human cancer remains controversial.

70.5.5  Reproductive and Developmental Toxicity Chlorpyrifos, as a highly lipophilic compound, is expected to cross the placenta and reach the fetus during pregnancy. Several studies, carried out with rats, mice, and rabbits, have focused on the potential embryotoxicity and teratogenicity of gestational exposure to CPF (Breslin et al., 1996; Deacon et al., 1980; Rubin et al., 1987). Available data suggest that CPF is not teratogenic and does not adversely affect reproduction. No consistent evidence of teratogenic effects was found in offspring when pregnant rats were fed doses as high as 15 mg CPF/kg/day for 10 days (Breslin et al., 1996; Rubin et al., 1987); there was also no evidence for fetotoxicity, measured as prenatal mortality or alterations in fetal weight. Signs of prenatal toxicity, fetal growth retardation, as well as increased incidence of malformations (i.e. minor skeletal variations) were observed only in the presence of frank maternal toxicity (Akhtar et al., 2006; Farag et al., 2003). Furthermore, several multigenerational studies have been performed in rats to investigate ChE inhibition, fetal development, and possible adverse effects on reproductive capacity, including fertility, following prenatal CPF ex­posure, giving negative results (Breslin et al., 1991, 1996). Only one study reported decreases in fetal body

Hayes’ Handbook of Pesticide Toxicology

weight at CPF doses (5 mg/kg/day), enough to give significant maternal toxicity and inhibition of brain ChE in dams (James et al., 1988). All the studies confirmed that RBC inhibition is the more sensitive marker for CPF-induced effects (Hoberman, 1998; Maurissen et al., 2000a). In humans, indoor and/or outdoor CPF exposure during pregnancy, estimated by using urinary levels of TCP in pregnant women or CPF levels in umbilical cord plasma as a direct biomarker for CPF in utero fetal exposure, has been associated with different birth outcomes, including decreased birth size (Whyatt et al., 2004), decreased gestational age at birth (Eskenazi et al., 2004), and decreased head circumference in the presence of low levels of PON1 in the mother (Berkowitz et al., 2004). In addition to the aforementioned developmental tox­ icity, other effects, including alterations in male reproductive end points and in the homeostasis of thyroid or sex steroid hormones, likely due to the impairment of their metabolism (Rose and Hodgson, 2005), have been associated with CPF exposure in animals and in humans (Eaton et al., 2008), where urinary TCP content was correlated with decreased testosterone levels (Meeker et al., 2006).

70.5.5.1  Developmental Neurotoxicity In the past years a growing body of literature has described the effects of CPF in the developing brain after gestational and early postnatal exposure leading to both neurochem­ical and behavioral alterations into adulthood. Adverse effects on the development of the nervous system have been reported following pre/perinatal CPF exposure both in rodents (Aldridge et al., 2005b; Carr et al., 2001; Ricceri et al., 2006; Richardson and Chambers, 2005; Venerosi et al., 2006, 2009; Zheng et al., 2000) and in humans (Rauh et al., 2006; Whyatt and Barr, 2001). After a single in utero CPF exposure, neonates showed higher susceptibility to brain ChE inhibition than dams (Chanda et al., 1995; Lassiter et al., 1998), whereas repeated exposure induced a higher degree of ChE inhibition in the maternal brain than in the fetus (Lassiter et al., 1998, 1999; Mattsson et al., 2000; Zheng et al., 2000). The age-related susceptibility to ChE inhibition could be ascribed to the greater ability of ChE to recover in the fetal brain than in the adult one, probably due to higher level of protein synthesis. The difference between the decrease of ChE in fetal and maternal brain has been shown to be dose-dependent (Lassiter et al., 1999; Richardson and Chambers, 2003). Indeed, the lower the CPF dosage (1–10 mg/kg/day, administered orally to dams), the higher the maternal/fetal ratio in brain ChE inhibition, likely due to efficient maternal metabol­ ism, which at low doses (1 mg/kg/day) prevents significant fetal exposure. On the other hand, although an in situ CPF bioactivation is possible in the fetus, catalyzed by CYP3A7 (Buratti et al., 2006), the contribution of fetal metabolism to the total CPF biotransformation is one to two orders of

Chapter | 70  Chlorpyrifos

magnitude lower than that catalyzed by 3A4 of the mother (Buratti et al., 2006). In addition, during pregnancy CYP3A7 is expressed also in the endometrium and placenta. Although its content per gram of tissue is only 0.6–5.5% of CYP3A7 content in the fetal liver, placental weight is about five times more than the fetal liver weight (Schuetz et al., 1993), indicating the importance of considering the mother and the fetoplacental unit as parts of the same system. This consideration was confirmed by recent data showing that the combined analysis of maternal (hair and blood) and infant (cord blood, infant hair, or meconium) matrices as optimum biomarkers to detect fetal exposure to environmental pesticides significantly increases the detection rate (Ostrea et al., 2009). However, at low CPF levels of exposure, the formation of the nontoxic metabolite is highly favored in the fetus, and since PON1 is expressed during prenatal periods at very low levels (Costa et al., 2005b), it is likely that the CYP-mediated detoxication reactions in the fetus exposed to CPF are the most efficient biotransformation pathway (Buratti et al., 2006). Although both AChE and ACh play a key role in brain morphogenic development (Brimijoin and Koenigsberger, 1999; Lauder and Schambra, 1999; Layer and Willbold, 1995), CPF-induced alteration in brain development, causing neurochemical and behavioral effects, has not been simply related to cholinergic mechanisms. Indeed, there is a high degree of variability (from no/marginal to substantial) in the inhibition of fetal or neonatal brain AChE activity in the pups, with effects persisting to adulthood, and many mechanisms may be at work (Slotkin et al., 2006). Cognitive and behavioral effects in the pups, as well as general signs of cholinergic toxicity, have been evidenced in the presence of significant brain ChE inhibition (70%) (Clegg and van Gemert, 1999; Zhao et al., 2006). The inhibitory effect on brain ChE due to exposure to high CPF levels during gestation and/or lactation persisted up to a few days after the cessation of the postnatal treatment (PND9) (Richardson and Chambers, 2003) or during lactation (Mattsson et al., 2000). Many of the CPF neurobehavioral effects were found to be time- and sex-dependent, with females or males affected to a greater extent when CPF exposure was pre- or postnatal, respectively (Dam et al., 2000; Levin et al., 2001, 2002). Prenatal CPF exposure in rats, at a subtoxic dose (i.e., 1 mg/kg/day), produced in females long-lasting behavioral impairment (up to adulthood), such as locomotor hyperactivity and working and reference memory deficits (Levin et al., 2001). After postnatal CPF exposure in females, a masculinization of behavior was evidenced (i.e., reduction in locomotion and decreased errors in the radial arm maze) (Dam et al., 2000). During the neonatal period critical for brain development, even a rapid and transient ChE inhibition was able to induce a cascade of events, ending in gender-dependent persistent changes in behavioral performance (Dam et al., 2000). Moreover, following a postnatal acute (on PND17) (Moser and Padilla, 1998) or repeated gestational ­subcutaneous (gestational days 12–19) (Chanda and Pope,

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1996) CPF injection in rats, pups showed a decrease in total muscarinic ACh receptor binding, along with a deep inhib­ ition of brain and blood ChE, and behavioral clinical signs in the early postnatal days. The occurrence of cognitive impairment also in the absence of a significant inhibition of fetal brain ChE indicates that at low exposure levels noncholinergic mechanisms may be relatively more important. It is now clear that CPF influences different basic processes of neural cell development, eliciting widespread disruption of neural cell replication, DNA synthesis, gene transcription, and cell differentiation and impairment in axonogenesis, synaptogenesis, and synaptic function (Barone et al., 2000; Pope, 1999; Slotkin, 1999, 2004). Most of these effects were evidenced at CPF doses below the threshold inhibition of brain ChE and were associated with persistent neurobehavioral deficits (Levin et al., 2001, 2002). In addition, several rat studies have shown that CPF altered neuronal development related to serotonin (5HT) and dopamine (Aldridge et al., 2003; Slotkin et al., 2002), hypothesized as partially responsible for CPF-induced longterm behavioral changes (Aldridge et al., 2005a,c). However, despite the number of suggestions for a mode of action for CPF alternative to its effects on the cholinergic system, there is no conclusive evidence for the identification of the relevant targets and specifically for the understanding of whether the observed neurochemical changes (i.e., serotonin, adenylyl cyclase system, differential gene expression) are primary effects, distinct from AChE inhibition. As concluded by the EPA (U.S. EPA, 2008b), the inability to discern is usually due to study design issues (different route and window of exposure, species, way to assess behavioral effects), lack of time course information and/or toxicokin­ etic data on tissue dosimetry, and the high degree of variability in responses, resulting in limited differences between treated and control groups, making it difficult to distinguish spontaneous variability from treatment-related effects. In addition, at present specific mechanistic in vitro data showing, for example, whether a direct interaction of CPF with the serotonin transporter or its receptors can occur, are not available. However, when taken together they provide a weight of evidence with respect to CPF-induced persistent effects on neurodevelopment. In addition to the wide number of animal studies, several human studies investigated the possible association between CPF exposure during pregnancy and neurodevelopmental perturbation, with contrasting results. In particular, three major prospective epidemiologic cohort studies have been carried out to look at pre and postnatal pesticide exposure in mothers and infants, birth outcomes, and long-term childhood neurobehavioral and neurodevelopment outcomes (U.S. EPA, 2008c). Some data suggest that prenatal CPF exposure, measured as CPF levels in maternal and umbilical cord blood, may be related to delays in cognition and psychomotor function in children up to 36 months (Rauh et al., 2006). On the other hand, no significant association with any changes in cognitive

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and behavioral outcomes was evidenced in children prenatally exposed to CPF, by using urine TCP levels as marker of ex­posure (Eskenazi et al., 2007). The weight of the evidence led the EPA to conclude that exposure to OPTs at sufficiently high levels during gestation, particularly in susceptible populations, may result in neurodevelopmental outcomes. At lower exposure levels, as those found in food, such effects have not been observed (U.S. EPA, 2008c). However, the potential for multiple OPT exposure impacting the effects observed in children cannot be ruled out (U.S. EPA, 2008c).

70.6  Toxic interactions Since different pesticides can be concurrently used in agricultural practices, simultaneous or sequential exposure is possible among agriculture workers and general population consuming contaminated foods. When combined exposure to pesticides sharing common biotransformation pathways and/or mechanism of action occurs, such as in the case of OPTs having in AChE the same target, adverse effects may be the result of toxicological interactions. As requested by the FQPA (FQPA, 1996), in the United States the cumulative risk associated with exposure to different OPTs or mixed formulations is considered. Based on the common mech­ anism of action, response additivity is usually assumed. The relative potency factor (RPF) method for dose addition has been proposed for the group evaluation of OPTs (U.S. EPA, 2006). The relative contribution within a given mixture is cal­culated with respect to the inhibition power of methamidophos, used as the index chemical. The response of the mixture is then predicted as the sum of equivalent exposures. In this respect, a PBPK/PD model for CPF and diazinon and their metabolites is available for rats. It showed that there are no differences in the pharmacokinetics of either the parent compounds or their respective metabolites in the case of binary mixtures, and ChE inhibition can be described using the dose-additive model (Timchalk and Poet, 2008). However, although OPT pesticides share a qualitatively similar, but quite often quantitatively different, extent of metabolism, they present a very complex pattern of interactions, which in many cases does not simply correspond to additivity. In vitro data are available showing that mixtures of CPFO and azinphos methyl-oxon, at high concentrations, induced greater than additive effects on ChE activity in the blood and brain of rats (Richardson et al., 2001). The occurrence of in vivo interactions following exposure to CPF and parathion has also been demonstrated in both adults and neonatal rats, showing that the sequence of exposure could markedly impact the toxicity, very likely due to the depletion of CE involved in the detoxification reaction or the differential efficiency of PON-1 toward CPFO and paraoxon (Kacham et al., 2006; Karanth et al., 2001, 2004). More extensive toxicity, in terms of lethality, clinical signs, and ChE inhibition, was evidenced in rats pre-exposed to CPF

Hayes’ Handbook of Pesticide Toxicology

followed by parathion compared with rats exposed simultaneously to the two OPTs or pre-exposed to parathion. The mechanism to explain the interaction could be related to inhibition of paraoxon hepatic detoxification by CPF pretreatment. Indeed, CPF has been reported to inhibit CE at doses less than or equal to those inhibiting ChE (Buratti and Testai, 2005; Chambers and Carr, 1993; Chanda et al., 1997). CPFO is a much stronger CE inhibitor than CPF (Ki  22 nM and 7.5 M, respectively); therefore, an interaction can be anticipated in case of concomitant exposure with malathion, whose relatively low toxicity is due to the CEmediated efficient detoxification (Buratti and Testai, 2005). At relatively low malathion and CPF concentrations (i.e., 0.5–1 M), CPF itself can only partially inhibit the already active CE-mediated hydrolysis of malathion by direct interaction. However, CPF is efficiently metabolized to CPFO, thus limiting malathion detoxification. Although the sequestration of CPFO by CE works as a detoxification mechanism toward the CPF-induced effects, it can concurrently represent the cause for an increased malathion bioactivation to its oxon and consequently toxicity enhancement (Buratti and Testai, 2005). Another example is the CPFO inhibition of the initial step of the metabolism of permethrin (a pyrethroid pesticide). Permethrin is hydrolyzed in human microsomes and cytosol by esterase, which are efficiently inhibited by CPFO at concentrations as low as 60 nM with decreased permethrin detoxification (Choi et al., 2003, 2004). The administration to adult and preweanling rats of a mixture of five OPTs (CPF, diazinon, dimethoate, acephate, and malathion), in a ratio reflecting the relative dietary ex­posure estimates of the general population, resulted in a greater than additive response (synergism) observed at the lower doses of the mixture, corresponding to noneffective dose levels of each of the components. The predicted effective doses (ED20, ED50) for different endpoints, among them blood and brain ChE and behavior pattern, were about half those predicted by additivity. The potentiation was evidenced to a greater extent in young animals compared with adults (Moser et al., 2005). Detoxification factors and generally kinetic interaction have been suggested as the major causes for the observed synergy (Moser et al., 2005, 2006). Most OPTs share the same biotransformation pathway, being substrates of common CYPs, (Buratti et al., 2003; Sams et al., 2004; Tang et al., 2001) which could be the source for additional metabolic interactions. In a recent in vivo study CPF and diazinon were selected as a model for toxicokinetic-based interactions in rats (Timchalk and Poet, 2008). Concurrent exposure to CPF and diazinon produced a dose-dependent inhibition of plasma, RBC, and brain ChE activities, being mainly influenced by the presence of CPF. Exposure to high doses of the mixture resulted in the alteration of blood kinetics for both parent compounds, likely due to a competition between CPF and diazinon for the same CYP. This explanation is supported by in vitro data showing the ability of CPF to inhibit diazinon metabolism through competitive

Chapter | 70  Chlorpyrifos

or uncompetitive mechanisms (Wu et al., 2004). Indeed, it is well known that during CPF desulfuration highly reactive sulphur atoms are formed, binding irreversibly to specific CYPs (Halpert et al., 1980) and leading to a typical “suicidal” or mechanism-based inhibition. As a result, in the presence of CPF a reduction of the related CYP catalytic activity may occur with clear consequences for the metabolism of other compounds sharing the same CYPs for the biotransformation (Hodgson and Rose, 2007). As an example, since CYP2B6 is involved in both CPF bioactivation and the methyl hydroxylation of the carbamate insecticide carbaryl, a metabolic competition is reasonably expected between the two pesticides, as confirmed by an in vitro study with human liver microsomes (Tang et al., 2002). CPF has also been shown in vitro to be a potent inhibitor of the human metabolism of numerous pesticides belonging to different classes (Hodgson and Rose, 2006, 2007; Joo et al., 2007; Usmani et al., 2002). Metabolic interactions are not limited to other pesticides and have also been reported with endogenous substrates, such as testosterone and estradiol (Hodgson and Rose, 2007; Usmani et al., 2003, 2006). Interestingly, a clear in vitro interaction between different OPTs, including CPF and a clinical drug, the antidepressant imipramine, has been evidenced both in human liver microsomes and in recombinant CYPs (Di Consiglio et al., 2005). Imipramine demethylation to the pharmacologically active metabolite was significantly inhibited (up to 80%), starting at low pesticide concentrations (0.25–2.5 M). At concentrations close to the actual exposure levels, the inhibition appeared more efficient when CYP1A2 was active, consistent with the pivotal role of CYP1A2 in CPF bioactivation at low concentrations (Buratti et al., 2003). Therefore, exposure to OPTs in the micromolar range during medical treatment with imipramine may lead to a decreased drug bioactivation, resulting in high plasma levels of the parent compound with impairment of the pharmacological action and possible onset of ADRs. Chlorpyrifos may alter the serotonin (5HT) systems (Aldridge et al., 2005a,b), and indeed recent investigations have identified an interaction of CPF with terbutaline, a clinical drug used in the treatment of premature labor, which also produces impairment of the 5HT circuit (Aldridge et al., 2005c). Sequential exposure of rats first to terbutaline and then to CPF during the crucial period for neurodevelopment resulted in an increase of the adverse effects on 5HT function, with permanent consequences evident in juvenile and adult rats. No adequate epidemiological studies are available for interactions of CPF in mixtures. A few reports identified an association between combined exposure during pregnancy to CPF and other toxicants (i.e., other pesticides, like diazinon or environmental tobacco smoke) and fetal growth of newborns (Rauh et al., 2004, 2006). No evidence of interaction between environmental tobacco smoke and CPF exposure has been reported for the neurocognitive defects in infants. Although these data suggest that interactions between toxicants during

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pregnancy may occur, inducing adverse outcomes in infants and/or children, further investigations are needed to clarify the role of such interactions.

Conclusion CPF is, among the organophosphorothionates, one of most frequently used insecticides due to its effective, cost-competitive broad spectrum of activity when compared with alternative products. Although its toxicokinetic and acute toxicologic­al properties related to CYP-mediated oxon formation and AChE inibition have been known for a long time, hundreds of thousands of deaths per year still have been reported to occur in developing countries. Other aspects, including effects derived from long-term exposure, have not yet been completely clarified. In addition, the reported excess in children’s dietary intake has focused attention on CPF-induced neurodevelopmental effects possibly caused by perinatal exposure and on the occurrence of cognitive impairment. However, there is no conclusive evidence for the identification of the relevant targets, and it is still unclear whether these are primary effects, actually distinct from AChE inhibition. The possible interactions with other xenobiotics and the evaluation of cumulative risk are additional open issues.

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Chapter | 70  Chlorpyrifos

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organophosphorus and carbamate pesticides. (Office of Pesticide Programs, ed.), US EPA, Washington DC. U.S. EPA (2002). Interim Reregistration Eligibility Decision for Chlorpyrifos. US EPA, Washington, DC http://www.epa.gov/oppsrrd1/ reregistration/status.htm. U.S. EPA (2006). Organophosphorus Cumulative Risk Assessment, US EPA, Washington, DC. http://www.epa.gov/pesticides/cumulative/ 2006-op/index.htm. U.S. EPA (2008a). September 16–19, 2008: The Agency’s evaluation of the toxicity profile of chlorpyrifos. Appendix A: Metabolism. (US EPA, ed.), Washington, DC. http://www.epa.gov/scipoly/sap/meetings/2008/091608_mtg.htm. U.S. EPA (2008b). September 16-19,2008: The Agency’s evaluation of the toxicity profile of chlorpyrifos. Appendix C: Effects on the developing brain. (US EPA, ed.), Washington, DC. Available at: http:// www.epa.gov/scipoly/sap/meetings/2008/091608_mtg.htm. U.S. EPA (2008c). September 16–19, 2008: The Agency’s evaluation of the toxicity profile of chlorpyrifos. Appendix D: Human epidemiology: effects in children. (US EPA, ed.), Washington, DC. Available at: http://www.epa.gov/scipoly/sap/meetings/2008/091608_mtg.htm. U.S. Public Health Service (1995). Hazardous Substance Data Bank. 5-9. Washington DC. Usmani, K. A., Cho, T. M., Rose, R. L., and Hodgson, E. (2006). Inhibition of the human liver microsomal and human cytochrome P450 1A2 and 3A4 metabolism of estradiol by deployment-related and other chemicals. Drug Metab. Disp. 34, 1606–1614. Usmani, K. A., Rose, R. L., Goldstein, J. A., Taylor, W. G., Brimfield, A. A., and Hodgson, E. (2002). In vitro human metabolism and interactions of repellent N,N-diethyl-m-toluamide. Drug Metab. Disp. 30, 289–294. Usmani, K. A., Rose, R. L., and Hodgson, E. (2003). Inhibition and activation of the human liver and human cytochrome P450 3A4 metabolism of testosterone by deployment-related chemicals. Drug Metab. Disp. 31, 384–391. Valcke, M., and Bouchard, M. (2009). Determination of no-observed effect level (NOEL)-biomarker equivalents to interpret biomonitoring data for organophosphorus pesticides in children. Environ. Health 8:5, http://www.ehjournal.net/content/8/1/5. van Gemert, M., Dourson, M., Moretto, A., and Atson, M. (2001). Use of human data for the derivation of a reference dose for chlorpyrifos. Regul. Toxicol. Pharmacol. 33, 110–116. Venerosi, A., Calamandrei, G., and Ricceri, L. (2006). A social recognition test for female mice reveals behavioral effects of developmental chlorpyrifos exposure. Neurotoxicol. Teratol. 28(4), 466–471. Venerosi, A., Ricceri, L., Scattoni, M. L., and Calamandrei, G. (2009). Prenatal chlorpyrifos exposure alters motor behavior and ultrasonic vocalization in CD-1 mouse pups. Environ. Health 8,1/12. http:// www.ehjournal.net/content/8/1/12. Waddell, B. L., Zahm, S. H., Baris, D., Weisenburger, D. D., Holmes, F., Burmeister, L. F., Cantor, K. P., and Blair, A. (2001). Agricultural use of organophosphate pesticides and the risk of non-Hodgkin’s lymphoma among male farmers (United States). Cancer Causes Control 12(6), 509–517. Warner, S. D., Gerbig, C. G., Strebing, R. J., and Molello, J. A. (1980). “Results of a Two-Year Toxicity and Oncogenicity Study of Chlorpyr­ifos

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Administered to CD-1 Mice in the Diet,” Dow Chemical Co., Midland MI. Watanabe, A. M. (1989). Cholinoreceptor-activating drugs. In “Basic and Clinical Pharmacology” (B. G. Katzung, ed.), pp. 70–105. Appleton & Lange, East Norwalk. Water, M. D., Simmon, V. F., Mitchell, A. D., Jorgenson, T. A., and Valencia, R. (1980). An overview of short-term tests for the mutagenic and carcinogenic potential of pesticides. J. Environ. Sci. Health B 15, 867–906. WHO (2007). WHO specifications and evaluations for public health pesticides: Chlorpyrifos. http://www.who.int/en/ Whyatt, R. M., and Barr, D. B. (2001). Measurement of organophosphate metabolites in postpartum meconium as potential biomarker of prenatal exposure: a validation study. Environ. Health Perspect. 109(4), 417–420. Whyatt, R. M., Rauh, V., Barr, D. B., Camann, D. E., Andrews, H. F., Garfinkel, R., Hoepner, L. A., Diaz, D., Dietrich, J., Reyes, A., Tang, D., Kinney, P. L., and Perera, F. P. (2004). Prenatal insecticide exposures and birth weight and length among an urban minority cohort. Environ. Health Perspect. 112(10), 1125–1132. Wilson, N. K., Chuang, J. C., Lyu, C., Menton, R., and Morgan, M. K. (2003). Aggregate exposures of nine preschool children to persistent organic pollutants at day care and at home. J. Expos. Anal. Environ. Epidemiol. 13, 187–202. Woodruff, R. C., Phillips, J. P., and Irwin, D. (1983). Pesticide-induced complete and partial chromosome loss in screens with repair-defective females of Drosophila melanogaster. Environ Mutagen 5, 835–846. Wu, H., Poet, T. S., and Timchalk, C. (2004). Inhibition of diazinon metabolism by chlorpyrifos in rat liver microsomes. Toxicol. Sci. 78(1-S), 425. Yano, B. L., Young, J. T., and Mattsson, J. L. (2000). Lack of carcinogenicity of chlorpyrifos insecticide in a high-dose, 2-year dietary toxicity study in Fischer 344 rats. Toxicol. Sci. 53, 135–144. Yasmashita, M., Tanaka, J., and Ando, Y. (1997). Human mortality in organophosphate poisoning. Vet. Human Toxicol. 39, 84–85. Yeary, R. A., Eaton, J., Gilmore, E., North, B., and Singell, J. (1993). A multiyear study of blood cholinesterase activity in urban pesticide applicators. J. Toxicol. Environ. Health 39(1), 11–25. Young, J. T., and Grandjean, M. (1988). Chlorpyrifos: 2-year dietary chronic toxicity-oncogenicity study in Fischer 344 rats. Study ID: K- 044793-079, dated December 23, 1988. [Dow; Submission 11462, reference 73. Reviewed by PMRA]. Lake Jackson Research Center, Texas, Dow Chemical Co. Zhang, Q.-Y., Dunbar, D., Ostrowska, A., Zeisloft, S., Yang, J., and Kaminski, L. S. (1999). Characterization of human small intestinal cytochromes P-450. Drug Metab. Disp. 27(7), 804–809. Zhang, X., Driver, J. H., Li, Y., Ross, J. K., and Krieger, R. I. (2008). Dialkylphosphates (DAPs) in fruits and vegetables may confound biomonitoring in organophosphorus insecticide exposure and risk assessment. J. Agric. Food Chem. 56(22), 10638–10645. Zhao, Q., Dourson, M., and Gadagbui, B. (2006). A review of the reference dose for chlorpyrifos. Regul. Toxicol. Pharmacol. 44(2), 111–124. Zheng, Q., Olivier, K., Won, Y. K., and Pope, C. N. (2000). Comparative cholinergic neurotoxicity of oral chlorpyrifos exposures in preweaning and adult rats. Toxicol. Sci. 55, 124–132.

Chapter 71

Malathion: A Review of Toxicology Inge M. Jensen1 and Paul Whatling2 1 2

Cheminova A/S, Lemvig, Denmark Cheminova, Inc., Arlington, Virginia

71.1  Introduction Malathion is an organophosphate insecticide with the chemical structure shown in Figure 71.1. Malathion is a broad-spectrum insecticide and has been successfully used worldwide for more than 50 years. It is primarily used as a plant protection product in agriculture for control of a wide range of insect pests on fruit and vegetable crops as well as for field crops, for protection of grain during storage, and particularly for control of the boll weevil in cotton, where it has successfully been employed in large-scale eradication programs. Malathion is also used in public health for control of vector-borne diseases such as malaria and West Nile virus. It is used in some areas of the world in animal husbandry for the control of ectoparasites on cattle, sheep, and poultry. Finally, malathion is used as a pharmaceutical product for control of head lice in humans. Humans may be potentially exposed to malathion in several ways. This may be during application of the pesticide, through consumption of treated food commodities, through contaminated drinking water, and through re-entering areas that have been treated with malathion. The toxicological properties of malathion relevant to all of these exposure

routes, which include exposures via oral, dermal, and inhalation routes, have been extensively studied and are discussed in this chapter. Malathion had been reviewed by major regulatory authorities throughout the world. The U.S. Environmental Protection Agency (EPA) reviewed the data on malathion as part of the implementation of the Food Quality Protection Act of 1996 and published a re-registration eligibility document (RED) in 2006 (U.S. EPA, 2006b). Malathion was also reviewed in the European Union as part of a general review of all plant protection products that were on the market at the time of implementation of a new harmonized legislation for regulation of pesticides. A conclusion report was published by the European Food Safety Authority (EFSA) in 2006. Malathion has been reviewed by the World Health Organization (WHO) in 1997 and 2003 and by the United Nations Food and Agricultural Organization (FAO) in 1999. These evaluations are published in the reports from the Joint Meeting on Pesticide Residues (JMPR, 2003). In 2003, the UK Committees on Mutagenicity and Carcinogenicity of Chemicals in Food, Consumer Products and the Environment (UK COM/COC) made a joint statement on the mutagenic and carcinogenic properties of malathion.

O

O

S

OC2 H 5

P

OC2 H 5

H3 CO H3 CO

S

O

OC2 H 5

H3 CO P H3 CO

OC2 H 5 S O

O

Malathion

Malaoxon

Figure 71.1  Formation of malaoxon. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

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These official evaluations were mainly based on unpublished data and information submitted by the manufacturers but also on relevant data from the open literature. The data submitted by the manufacturers are standard studies conducted according to internationally accepted regulatory guidelines, such as the Organization for Economic Co-operation and Development guidelines for testing of chemicals and the Office of Prevention, Pesticides and Toxic Substances guidelines for testing of pesticide chemicals, and conducted according to internationally accepted principles for Good Laboratory Practice. This chapter provides a summary of the official evaluations conducted by regulatory authorities in recent years and therefore contains previously unpublished data. The individual evaluations are publicly available and details can be found in the reference list.

71.2  Mode of action Malathion is a nonsystemic acaricide and insecticide of low mammalian toxicity with contact, stomach, and respiratory action. The residual effect is limited because the compound is degraded relatively quickly in the environment. The half-life in soil, water, and plants is less than 1 week. The residual effect can be prolonged by various formulation technologies. Malathion is a pro-insecticide that is bioactivated by oxidative desulfuration to the cholinesterase inhibitor (ChEI) malaoxon following absorption by the target organism (see Figure 71.1). In mammals, malathion undergoes metabolic activation (oxidation) to form malaoxon, which can inhibit chol­inesterase enzymes in blood, brain, and nervous tissue throughout the body. The amount of malaoxon formed from malathion in mammals in vivo is small. This, combined with its rapid metabolism by detoxifying enzymes (carboxyesterases and glutathione transferases), explain its low mammalian toxicity.

71.3  Basic physical and chemical properties Malathion is a colorless to pale yellow-colored liquid at room temperature with a characteristic odor and a relative density of 1.23 g/ml at 20°C. It has a freezing point of less than 220°C and starts to decompose at 174°C. With a low vapor pressure (4.5  104 Pa at 25°C) and relatively high aqueous solubility (148.2 mg/l in water at 25°C), malathion is classified as relatively nonvolatile (Henry’s law constant  1.01  103 Pa m3/mol). Malathion is relatively lipophilic as characterized by its octanol–water partition coefficient (log Pow2.748). The rate of hydrolysis of malathion increases with pH. At pH 9, the hydrolysis half-life of malathion is 11.8 h at 25°C, whereas at pH 5 the half-life is

Hayes’ Handbook of Pesticide Toxicology

107 days. Aqueous photolysis had to be conducted at pH 4 to minimize hydrolysis. The half-life of malathion for photolysis is 93 days (sensitized) and 156 days (nonsensitized). Malathion has a flash point of 173  2°C and is not explosive under normal conditions of storage and use. It does not have oxidizing or reducing properties.

71.4  Impurities Highly purified malathion has very low toxicity to mammals. However, all technical grades of malathion, as well as all malathion end-use products, will inevitably contain various impurities potentially leading to higher toxicity of the material. The majority of the impurities are formed during the manufacturing process, but storage conditions may affect the composition of the material. Because malathion has been used for such a long period of time, and therefore different manufacturing processes have been used, a range of impurities have been reported in malathion technical and formulated products from different sources. Some of the impurities identified are more toxic than malathion. Some impurities may potentiate malathion toxicity by inhibiting detoxifying enzymes. Others exert their toxic effect via mechanisms apart from acetyl cholinesterase (AChE) enzyme inhibition. In addition, some of the impurities are known to interfere with each other via synergistic or antagonistic effects. When malathion is produced according to modern, tightly controlled manufacturing processes, the toxicologically relevant impurities in technical material are malaoxon, MeOOSPS-triester, MeOOOPS-triester, and isomalathion (FAO/WHO, 2003). Of these, the first three are formed during the manufacturing process, whereas the latter is primarily formed during storage (as a function of time and temperature). The toxicity of any grade of malathion is therefore highly dependent on, and cannot be predicted without, knowledge of the impurity pattern of the material. This is of fundamental importance and the International Union of Pure and Applied Chemistry has published a technical report on the significance of impurities in crop protection products (Ambrus et al., 2003). The authors of this report comprised experts from government agencies, academia, and industry. A key recommendation of this report was that detailed information on composition of technical material and formulated products together with appropriate toxicological studies should be provided to regulatory authorities. Moreover, for toxicologically relevant impurities, maximum permissible concentrations should be specified. As a consequence, information on overall purity and also detailed information on the individual impurities present in the test batch are essential when evaluating the result of toxicity testing on malathion.

Chapter | 71  Malathion: A Review of Toxicology

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O

O

S

OC2 H 5

P

OC2 H 5

H3 CO H3 CO

S

O

OC2 H 5

H 3 CS P

OC2 H 5

H3 CO

S O

O

Malathion

Isomalathion

Figure 71.2  Formation of isomalathion.

O

S

OH

H3 CO P H3 CO

OC2 H 5 S O

O

S

O

OC2 H 5

H3 CO P H3 CO

OH S

S

OH

H3 CO P H3 CO

OH S

O

MMCA alpha- and beta isomers

O

MDCA

Figure 71.3  Structures of MMCA and MDCA.

The impurity isomalathion, which is formed by isomerization of malathion (Figure 71.2), is a special concern because of its potentiating effect. The rate of formation is dependent on time and temperature so that prolonged storage at elevated temperature will lead to increased isomalathion formation. The effect of isomalathion on malathion toxicity is discussed in detail later.

71.5  Metabolism of malathion Malathion is rapidly metabolized in mammals as well as in plants and the environment. Details of rates and routes of metabolism have been published in a review (Roberts and

Hutson, 1999), with considerably more unpublished information from regulatory reviews cited in the text. The metabolic pathways in mammals, plants, and the environment are broadly similar, with a major metabolic process being hydrolysis of ester bonds of malathion to form malathion monocarboxylic acid (MMCA) and malathion dicarboxylic acid (MDCA), as shown in Figure 71.3. These acids are further degraded to low-molecular-weight compounds including some that occur naturally. A relatively minor route, but important in mode of action and toxicity, is oxidation to malaoxon. A study conducted in Sprague–Dawley rats showed that malathion is rapidly absorbed, biotransformed, and rapidly eliminated from the body. Elimination is via excretion,

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O O S O

O

H3 CO P H3 CO

OC2 H 5

H3 CO SH

P H3 CO

OC2 H 5

OC2 H 5

H3 CO P H3 CO

OH S

S

O O

O

Malaoxon O

S

OH

H3 CO P S

S

H3 CO

OC2 H 5

H3 CO

OC2 H 5 S

H3 CO P

H3 CO

P SH

H3 CO

OC2 H 5

O

S

MMCA O

Malathion O

O

O S

OC2 H 5 OC2 H 5 HO

S

OC2 H 5

H3 CO

P

OC2 H 5

H3 CO

H3 CO

OH

P

HO

S

OH S O

O

O

Desmethyl Malathion (DMM)

MDCA

Figure 71.4  Metabolic pathways for malathion in the rat.

mainly in urine and to some extent in feces as mono- and dicarboxylic acids and other metabolites. More details on the absorption, distribution, metabolism, and excretion (ADME) studies on malathion are given later. Proposed metabolic pathways for malathion in the rat are shown in Figure 71.4. The metabolism of malathion in plants has been studied in several crops (i.e., cotton, lettuce, alfalfa, and wheat). Malathion is extensively metabolized in plants, leading to the incorporation of radiolabeled carbon into natural plant constituents. Unchanged malathion was found in each crop matrix tested. Metabolites identified but present at less than 10% of the total radioactive residue (TRR) include malaoxon, isomalathion, diethyl maleate, monoethyl maleate, diethyl mercaptosuccinate, MDCA, MMCA, diethyl methylthiosuccinate, diethyl fumarate, desmethyl malathion, and tetraethyl dithiodisuccinate. In one case, MMCA was present at 12.8% of the TRR in lettuce. Some of the radioactive residues were characterized as organic acids and sugars, demonstrating incorporation of radioactivity

into the tricarboxylic acid (TCA) cycle. Radioactivity was also incorporated into other endogenous plant constituents, such as cell wall fractions including starch, protein, pectin, lignin, hemicellulose, and cellulose. The metabolic pathway in plants operates via deesterification of parent malathion to form MDCA. This compound is subsequently cleaved to yield succinic acid. The succinate may then be incorporated into other small organic acids and sugars via the TCA cycle. The radioactivity is then extensively incorporated into natural plant constituents (Roberts and Hutson, 1999). The toxicology of malaoxon has been studied because malaoxon is more toxic than malathion. However, because other metabolites are less toxic than malathion, the relative toxicity for these metabolites has been calculated from the ratio of red blood cell (RBC) cholinergic response values for each metabolite and for malathion. The results of these investigations provided estimated relative potency factors for RBC cholinesterase inhibition for desmethylmalathion, MMCA, and MDCA, as summarized in Table 71.1.

Chapter | 71  Malathion: A Review of Toxicology

Table 71.1  Relative Potency Factors for Malathion Metabolites Metabolite

Relative Potency Factor

Malaoxon

22

Desmethyl-malathion

0.41

Malathion monocarboxylic acid

0.43

Malathion dicarboxylic acid

0.12

71.6  Human exposure to malathion Bearing in mind the varied uses of malathion, there are a number of potential routes of exposure to humans. For agricultural use, exposure can potentially occur during application at various stages, namely during mixing of the formulation and loading of the application equipment. Clearly, these exposures can be kept to a minimum by the sensible use of protective clothing and by following label instructions. Baker et al. (1978) reported more than 2800 cases of poisoning in Pakistan in July 1976 due to the use of malathion products (three different preparations were used) in a malaria prevention program. RBC cholinesterase was depressed at an average maximum of 45% using a brand that contained the highest amount of isomalathion. Cholinesterase levels were significantly lowered among workers who complained about headaches, blurred vision, and vomiting. Overall, the depression in cholinesterase activities correlated well with the content of isomalathion in the three products used. In vivo animal studies showed a correlation between chol­ inesterase activity depression and isomalathion content in the products. Improper work practices, leading to extensive dermal exposure, were observed frequently during the study. The authors concluded that the extent and severity of malathion intoxication in the study was caused by negligent and improper use of the products during mixing, loading, and spraying and by the fact that two of the three products contained increased concentrations of isomalathion and other products more toxic than malathion. Exposure to malathion can also potentially occur during re-entry of treated areas. The potential exposure to malathion residues in treated fields has been extensively investigated in studies conducted in the United States by the Agricultural Re-entry and Outdoor Re-entry Exposure Task Forces. Malathion was shown to rapidly degrade on a variety of fruit, vegetable, and leaf surfaces, as well as on turf. For re-entry into malathion-treated agricultural fields, the U.S. EPA has calculated safe re-entry periods of 12 h to 3 days, depending on the crop and activity (U.S. EPA, 2006a). Bystanders and those living near agricultural fields where pesticides are applied could also potentially be exposed to some extent via drift. The U.S. EPA considered this exposure

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route in its risk assessment and concluded that there is no concern (U.S. EPA, 2006a). Exposure by this route can be minimized by applying the chemical in appropriate weather conditions according to Good Agricultural Practices. Exposure by consumption of treated crops and/or drinking water is possible. Data on residues in edible commodities after treatment according to typical use practice have been used to set maximum residue levels, or tolerances, for each crop. Potential concentrations in drinking water can be estimated using models as well as by using available monitoring data. All of this information can be used to calculate a total potential dietary (food  drinking water) exposure of malathion and any relevant metabolites. The potential risk to consumers via this route has been thoroughly evaluated by regulatory authorities such as the U.S. EPA (2006) and the European Food Safety Authority (EFSA, 2006), as well as by the Joint FAO/WHO Meeting on Pesticide Residues (1999). Humans may also be exposed by direct application of malathion for head lice control, where malathion is applied as a low-concentration shampoo or lotion. Humans may also be potentially exposed during malaria control programs and other vector control programs.

71.7  Effects seen in animal toxicology studies The following sections summarize details of the wide range of metabolism and toxicology studies with malathion.

71.7.1  Absorption, Distribution, Metabolism, and Excretion Malathion is rapidly absorbed, biotransformed, and then quickly eliminated from the body. Elimination is via excretion mainly in urine and to some extent in feces primarily as mono- and dicarboxylic acids of malathion. Malathion or its metabolites do not accumulate in the tissues. As an example, an ADME study with rats is summarized next. Following low (40 mg/kg BW) or high (800 mg/kg BW) oral dosing of 14C malathion, elimination of 14C was rapid and extensive in rats of both sexes. Approximately 76–88% of the administered dose was excreted in urine and 7–11% in the feces. In the low-dose group, approximately 80% of the dose was eliminated in urine within 12 h. High-dose males excreted only approximately 48% of the total dose at this time; high-dose females excreted approximately 67%. At both dose levels, cumulative excretion through both urine and feces (0–72 h) accounted for greater than 90% of the total initial dose in both sexes. Pretreatment with nonlabeled malathion did not affect the excretion profile. Less than 1% of the administered dose was recovered in blood and tissues. Among the tissues, liver contained the most radioactivity, followed by skin, fat, and the gastrointestinal tract.

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The major metabolites in urine (comprising 80% of the total radioactivity) were - and -MMCA and MDCA. The remaining radioactivity was distributed among seven other metabolites: desmethyl malathion, O,O-dimethyl phosphorothioic acid, fumaric acid, 2-mercapto-succinic acid, O,O-dimethyl phosphorodiothioic acid, monoethyl fumarate, and malaoxon. Importantly, malaoxon was observed only in urine samples, accounting for less than 2% of the total radioactivity. The metabolic profile of fecal metabolites was comparable to that in urine samples. The hydrolysis of the carboxylester moiety of malathion by tissue (possibly liver) or that by plasma carb­ oxylesterases, resulting in - and -MMCA or MDCA, are the major pathways of metabolism. It is postulated that the malaoxon is formed by oxidative desulfuration of malathion by hepatic microsomal enzymes, and malaoxon is further metabolized by phosphatases. Hydrolysis by phosphatases would yield O,O-dimethyl phosphorothioic acid (from malaoxon) and O,O-dimethyl phosphorodithioic acid (from malathion). Other metabolic pathways include dealkylation probably by glutathione-S-transferases. Glutathione-dependent demethylation yields S-methyl glutathione and the corresponding desmethyl phosphate compound. A proposed metabolic pathway for malathion in the rat is shown in Figure 71.4.

71.7.2  Review of Toxicology Studies Because the toxicology of pure malathion is very low but may be enhanced by the presence of certain impurities, it is essential to know the composition and impurity content of malathion used in toxicology studies in order to interpret the results appropriately (Ambrus et al., 2003). All the studies discussed here were conducted to standard regulatory guidelines extant at the time of the study with malathion of known composition. In the following sections, the nature of toxicology studies conducted with malathion and their findings are described. For each short-term, subchronic and chronic study, no-observed-adverse-effect levels (NOAELs) have been identified. A NOAEL is the highest dose tested at which no adverse effects were detected. The NOAEL is typically defined based on statistical comparisons of the response at each dose level with the control group but sometimes may also reflect subjective judgments on the biological significance of a particular effect. The lowest dose at which an effect is judged to be adverse is commonly referred to as the lowest-observed-adverse-effect level (LOAEL). The NOAEL/LOAEL approach has traditionally been used by regulatory agencies throughout the world to describe the results of toxicological studies and to establish safe exposure limits for risk assessment. However, it is recognized that this approach does not necessarily reflect the relationship between dose and response for a given chemical, nor does it reflect a uniform response. Specific limitations of the

Hayes’ Handbook of Pesticide Toxicology

NOAEL/LOAEL approach are well-known and have been discussed extensively (U.S. EPA, 2000). Benchmark dose (BMD) modeling has been proposed as an alternative to the traditional NOAEL/LOAEL approach for describing study results and for establishing safe exposure limits for risk assessment. A BMD is estimated by defining a response level of interest [e.g., a 20% difference from controls (BMD20)] and estimating the dose that results in that level of change based on a regression analysis of the response against dose. With this method, it is typical to also calculate a 95% lower confidence limit on the BMD estimate, abbreviated as the BMDL. The calculated BMDL value is commonly considered to be the equivalent to a NOAEL. The BMD approach provides significant advantages in dose–response analysis (U.S. EPA, 2000), particularly in comparing responses across different assays or across time points within the same assay. The advantages of the BMD approach include the following: (1) It is less sensitive to sample size compared to the NOAEL approach, (2) it is not as sensitive to dose selection, and (3) it accounts for the full shape of the dose–response curve. BMDs can often be calculated even when a NOAEL is not determinable. Use of the BMD approach has become routine by the U.S. EPA. Background information about the BMD model may be found at the U.S. EPA’s website. For the studies described here, in which cholinesterase activity has been measured, the use of BMD modeling has been incorporated as a means of describing the relationship between the malathion exposure and the cholinergic response. BMDs were calculated using the exponential dose–response model developed and currently in use by the U.S. EPA (U.S. EPA, 2006a). For malathion, RBC cholinesterase activity has been shown to be the most sensitive endpoint. Compared to many other organophosphates, the differential sensitivity of RBC to brain cholinesterase inhibition is larger for malathion. According to most international conventions, an inhibition of 20% for RBC cholinesterase is a reasonable value for establishing a point of departure for risk assessment (WHO, 1999).

71.7.3  Acute Toxicity Highly purified malathion has, as previously mentioned, very low toxicity. However, depending on impurity level, and especially the isomalathion content, acute toxicity can vary. Isomalathion is an inhibitor of AChE. In addition, it is a potent inhibitor of tissue and serum carboxylesterases, which are known to rapidly detoxify malathion. As a result, small concentrations of isomalathion (0.01–0.5%) present as an impurity in malathion products result in an increase in malathion’s acute toxicity. This is not a directly proportional effect; for example, the effect of 0.05% isomalathion is to decrease the acute oral LD50 of malathion 3-fold, but a

Chapter | 71  Malathion: A Review of Toxicology

further 10-fold increase in concentration to 0.5% isomalathion produces only a further halving of the LD50 (Umetsu et al., 1977). Similar results were obtained in another study comparing the LD50 of purified malathion with that of malathion containing added isomalathion in known quantities (Aldridge et al., 1979). The acute oral LD50 for purified malathion was 10,700 mg/kg BW. Increasing ratios of isomalathion to malathion markedly increased the toxicity of malathion by 12-fold for a 0.02 ratio and by 17-fold for a 0.04 ratio. Based on a recent study, commercially available technical-grade malathion is considered to be moderately toxic by the oral route (LD50 1778 mg/kg BW). The isomalathion content of the test batch was 0.44%. Malathion is not toxic via the dermal route (LD50  2000 mg/kg BW in the rat) or by inhalation (LC�50  5.2 mg/l air/4 h). It exhibits only slight and reversible irritation to rabbit skin and eyes. Malathion also tested negative for skin sensitizing in the mouse local lymph node assay and tested positive in the guinea pig maximization test. Clinical signs after administration of a single large oral dose of malathion include tremors, decreased locomotor activity, ataxia, sedation or lethargy, salivation, and respiratory depression. Administration of high doses of malathion through whole-body inhalation exposure caused effects such as abnormal breathing, excessive salivation, and ataxia. Some of these effects may be due to irritance caused by the test material. Surviving test animals usually recovered within 3–14 days after acute exposure.

71.7.4  Short-Term and Subchronic Toxicity Malathion toxicity was moderate to low in all short-term studies. The most sensitive endpoint in short-term studies is inhibition of AChE activity, with RBC AChE being more sensitive than brain AChE. Clinical signs of toxicity were in general mild. At higher dose levels, common effects are decreased food consumption and effects on body weight and body weight gain. Increased liver and kidney weights are seen in rat and dog studies.

71.7.4.1  Short-Term Oral Toxicity Short-term oral toxicity studies have been conducted in the rat and dog. In the 28-day dietary rat study with dose levels ranging from 50 to 20,000 ppm, the NOAEL for RBC cholinesterase inhibition was 500 ppm (52 mg/kg BW/day). In the 90-day rat study, the NOAEL was 500 ppm (34.4 mg/kg BW/day) based on toxicological-relevant RBC and brain cholinesterase inhibition at 5000 ppm. BMD modeling identified BMDL20 values of 22.1 and 39.3 mg/kg/day, respectively, for male and female rats.

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Other signs of toxicity seen at high(er) dose levels in the rat were effects on food consumption, reductions in body weight and body weight gain, increased liver weights accompanied by hepatocellular hypertrophy, increased kidney weights, changes in hematology (decreased hemoglobin, PCV, mean corpuscular volume, and mean corpuscular hemoglobin content and increased RBC counts), and changes in clinical chemistry (increased -glutamyl transpeptidase levels and decreased alkaline phosphatase and aspartate aminotransferase). In the 90-day study, the incidence of chronic nephropathy was similar between treated males and control males; however, the severity was greater at dose levels at and above 5000 ppm. In the 28-day dog study, the NOAEL for cholinesterase was 250 mg/kg BW/day following capsule-gavage dosing at 125–500 mg/kg BW/day. In the 1-year capsule-gavage dog study, slight decreases in food consumption and body weight were noted at the 250 mg/kg BW/day level. Other effects included decreases in RBC counts, hemoglobin concentration, and hematocrit and also sporadic clinical chemistry changes (decreased albumin levels, calcium levels, total protein, and A/G ratios and elevated lactate dehydrogenase and alkaline phosphatase). No NOAEL could be established due to inhibition of RBC cholinesterase at the lowest dose level (62.5 mg/kg BW/day). However, BMDL20’s for RBC cholinesterase inhibition were calculated to be 38.6 and 42.9 mg/kg/day for males and females, respectively.

71.7.4.2  Dermal Toxicity The dermal toxicity of malathion has been examined in two 21-day toxicity studies using the rabbit. In the first 21-day dermal toxicity study (1989), malathion was applied at dose levels ranging from 50 to 1000 mg/kg BW/day (exposed for 6 h a day, 5 days a week). Application of the test item caused slight irritation (edema and erythema) of rabbit skin. Reduction in cholinesterase activity was the only treatmentrelated effect. The NOAEL was 50 mg/kg BW/day based on inhibition of RBC cholinesterase observed at the next highest dose level (300 mg/kg/day). Benchmark dose analysis was performed using the cholinesterase data from this study and the results are presented in Table 71.2. Results of this analysis demonstrated that RBC cholinesterase was much more sensitive to malathion exposure compared to brain cholinesterase, and males were more sensitive than females. The lowest BMDL value was 135 mg/kg/day for RBC ChEI in males; this value is considered to be equivalent to a study NOAEL. In the second 21-day dermal toxicity study (2006), malathion was applied at dose levels ranging from 75 to 500 mg/kg/day (daily exposures for 6 h/day over a 21-day period). Dermal irritancy was observed in all dosed groups (except control), and the severity increased with dose. There were no other treatment-related observations except

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Table 71.2  Benchmark Dose Modeling Results: 21-Day Dermal Toxicity Study (1989) Compartment

Sex

BMD (mg/ kg/day)

BMDL (mg/ kg/day)

BMD20, RBC ChEI

Male

168

135

Female

178

143

BMD10, ChEI in cerebellum

Male

698

680

Female

695

677

BMD10, ChEI in cerebrum

Male

295

279

Female

330

300

Table 71.4  Meta-analysis of RBC Cholinesterase Data from the 1989 and 2006 21-Day Dermal Toxicity Studies Compartment

Sex

BMD (mg/ kg/day)

BMDL (mg/ kg/day)

BMD20, RBC ChEI

Male

211

149

Female

179

127

for a reduction in cholinesterase activity. The NOAEL for this study was determined to be 100 mg/kg BW/day based on RBC cholinesterase inhibition observed at the next highest tested dose level of 150 mg/kg/day. Benchmark dose analysis was performed using the cholinesterase data from this study and the results are presented in Table 71.3. Results of this analysis demonstrated that RBC cholinesterase was more sensitive to malathion exposure compared to brain cholinesterase. In this study, females appeared to be slightly more sensitive to malathion exposure. The lowest BMDL value for the most sensitive ChE compartment (RBC ChE) was 127 mg/kg/day in female rabbits; this value is considered to be equivalent to a study NOAEL. A meta-analysis of the RBC ChE data from both dermal toxicity studies was conducted using BMD modeling to identify an overall point of departure for risk assessment (Table 71.4). The point of departure (BMDL20 for RBC ChEI) was determined to be 127 mg/kg BW/day.

skin/fur is very likely; the studies do not therefore demonstrate effects after only inhalation exposure. In the 14-day study with dose levels ranging from 0.56 to 4.28 mg/l, RBC cholinesterase activity was inhibited at all dose levels (0.56–4.23 mg/l, 6 h a day, 5 days a week). In the 13-week whole-body inhalation study in rats with dose rates ranging from 0.1 to 2.0 mg/l, treatment-related clinical signs were seen at 0.45 and 2.0 mg/l. Increases in absolute and relative liver and kidney weights were also seen at these dose levels. Lungs of the high-dose animals were generally heavier than those of the controls; however, no histopathological correlation was evident. The NOAEL was 0.1 mg/l based on inhibition of RBC AChE levels, whereas the NOAEL for inhibition of brain AChE was 0.45 mg/l. Benchmark dose analyses have been performed on the cholinesterase data from the two inhalation toxicity studies. For the 14-day study, the lowest BMDL20 was determined to be 0.21 mg/kg/day for RBC cholinesterase inhibition in females; this is considered to be equivalent to a study NOAEL. For the 13-week study, the lowest BMDL20 value was determined to be 0.098 mg/kg/day for RBC cholinesterase inhibition in females. This analysis supports the conclusion that the study NOAEL based on cholinesterase inhibition is approximately 0.1 mg/kg/day. In addition to the effects on cholinesterase activities, malathion treatment had an irritating effect on the nasal and laryngeal mucosa. Exposure-related histopathological findings were observed at all dose levels and included a high incidence of laryngeal hyperplasia and a dose-related incidence of degeneration and/or hyperplasia of the olfactory epithelium of the nasal cavity. The severity of these findings ranged from slight to moderate in the exposure groups. Rodent nasal mucosa contains high levels of xenobiotic-metabolizing enzymes, including carboxylesterases. It is postulated that locally generated mono- and dicarboxylic acids of malathion could be responsible for the observed damage and irritation to the mucosa cells.

71.7.4.3  Inhalation Toxicity

71.7.5  Genotoxicity

Two inhalation toxicity studies in rats have been conducted (14 days and 13 weeks), both with whole-body exposure. With whole-body exposure, oral exposure after licking the

The genotoxicity of malathion has been investigated in in vitro and in vivo studies. The design and outcome of these studies are described here.

Table 71.3  Benchmark Dose Modeling Results: 21-Day Dermal Toxicity Study (2006) Compartment

Sex

BMD (mg/ kg/day)

BMDL (mg/ kg/day)

BMD20, RBC ChEI

Male

211

149

Female

179

127

BMD10, ChEI— whole brain

Male

221

116

Female

206

129

Chapter | 71  Malathion: A Review of Toxicology

The results of the Ames test with Salmonella typhimur­ ium and Escherichia coli and the in vitro test for unscheduled DNA synthesis (UDS) with rat primary hepatocytes were negative. However, an in vitro gene mutation test using L5178Y TK/ mouse lymphoma cells, dose levels 125–2200 g/ml, gave a positive result. Malathion caused a dose-related response on cloning efficacy, growth rate, and mutation frequency, but findings were only statistically significant in the upper dose range (1800 g/ml) and were associated with marked cytotoxicity. Malathion tested positive for clastogenicity in the in vitro chromosome aberration test with human lymphocytes (at dose levels of 100–1200 g/ml). There was a significant increase in chromosome aberrations at a dose level of 450 g/ml but only in the absence of metabolic activation. This finding was associated with moderate cytotoxicity. Malathion tested negative in an in vivo chromosome aberration test in rat bone marrow cells carried out at dose levels up to 2000 mg/kg BW/day. In an in vivo DNA repair assay in male rat hepatocytes (in vivo UDS), rats received malathion at dose levels of 500, 1000, and 2000 mg/kg BW. Each dose level was tested with two different sampling points: 2–4 and 12–16 h. There was no increase in net nuclear grain counts at any dose level or at any sampling point; that is, malathion was not genotoxic in this assay. Overall, the results of these studies show that malathion has a potential for genotoxicity when tested in vitro at cytotoxic dose levels, but in vivo studies are negative. Consequently, malathion is not considered to be mutagenic in humans (for references see EFSA, 2004, 2006; UK COM/COC, 2003; U.S. EPA, 2006b).

71.7.6  Long-Term Studies Throughout the years, the toxicity of malathion has been thoroughly investigated in numerous long-term exposure studies using various animal species. Two guideline studies are currently considered by regulatory authorities throughout the world to be the most relevant – an 18-month carcinogenicity study in mice and a 2-year chronic toxicity/ carcinogenicity study in the rat. In both of these studies, the inhibition of AChE activity was determined to be the most sensitive effect. Thus, the use of AChE activity for setting reference doses for risk assessment is generally considered by regulatory agencies to be protective of all other toxicological effects that may occur at higher dose levels. A summary of the two critical studies is provided next. In the 18-month carcinogenicity study, technical-grade malathion was administered in the diet to groups of 65 male and 65 female mice at dose levels ranging from 100 to 16,000 ppm. Mortality rates, clinical signs of toxicity, and hematological parameters were not affected by treatment with malathion at any dose. At 16,000 and 8000 ppm in both males and females, treatment-related effects

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included decreased absolute body weights, decreased food consumption, and decreased RBC and brain cholinesterase levels. Based on the reductions in feed consumption and body weights, it was concluded that the maximum tolerated dose (MTD) was exceeded in males at 8000 and in both sexes at 16,000 ppm. At these same dose levels, there was significant liver toxicity, characterized by increased liver weight, hepatocellular hypertrophy, and a treatment-related increased incidence of hepatocellular adenomas observed in both male and female mice. There was no treatmentrelated increase in the incidence of liver carcinomas. Because the increased incidence of hepatocellular adenomas was associated with significant liver toxicity at dose levels exceeding the MTD, it is considered most likely that they were induced through a nongenotoxic mechanism and not significant to humans. Nasal toxicity characterized by degeneration and inflammation of the olfactory epithelium occurred at and above 800 ppm. The NOAEL for systemic effects was determined to be 800 ppm (143 mg/kg/day in males and 167 mg/kg/day in females) based on the effects seen at the LOAEL of 8000 ppm (1476 and 1707 mg/kg/day in males and females, respectively). The NOAEL for cholinesterase inhibition was 100 ppm (17.4 mg/kg/day) based on inhibition of RBC cholinesterase at higher dose levels. In the 2-year chronic toxicity and oncogenicity study, rats were exposed to malathion in the diet for 24 months at dose levels ranging from 100/50 to 12,000 ppm. Acetylcholinesterase inhibition, increased mortality, and decreased body weight and feed consumption were noted in the two high-dose groups of 6000 and 12,000 ppm. The MTD was exceeded in males at the two top dose levels and in females at 12,000 ppm. In the kidneys, chronic inflammation/nephropathy was observed with greater severity in the two high-dose groups. This finding correlated with surface irregularities of the kidneys that were noted macroscopically and with increased kidney weights. Degeneration and focal hyperplasia of the nasal epithelium and statistically significantly increased liver weights were seen in the two high-dose groups. The incidence of liver adenomas was statistically significantly increased in the females of the high-dose group. The liver adenomas are likely a development from hypertrophy and hyperplasia after prolonged strain on the detoxification activities of the liver and/or an effect of treatment on the background pathology of aging rats; this effect is not considered to be relevant to humans. A total of four nasal tumors were identified (one in males and one in females in each of the two highest dose levels): three nasal respiratory adenomas and one olfactory epithelial adenoma. Nasal adenomas are rare neoplasms, but all four occurred in the two exposure groups showing considerable nasal toxicity (including epithelial degeneration, hyperplasia, and inflammation). The most plausible explanation for the observed nasal irritation is prolonged high-level exposure of nasal epithelium to malathion from

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food either as a vapor or absorbed to inhaled food particles. In the nasal epithelium, malathion would be metabolized into mono- and dicarboxylic acids, and exposure to these acids would lead to local irritancy and cytotoxicity. This condition produces a state of reactive hyperplasia, one of the major causative factors in tumors. The overall NOAEL of the study was 500 ppm (29 mg/ kg BW/day) based on RBC and brain cholinesterase inhibition at higher dose levels. BMD modeling has been performed using the RBC and brain cholinesterase data from this study. A summary of the meta-analysis of data obtained for all measurement time points in the study is provided in Table 71.5. The modeling demonstrated that RBC cholinesterase inhibition was much more sensitive compared to brain cholinesterase inhibition, and that males may be slightly more sensitive to malathion exposure compared to females. The lowest BMDL20 value was 52 mg/kg/day for male RBC cholinesterase inhibition; this value is considered to be equivalent to a study NOAEL. Use of this BMDL20 of 52 mg/kg/day for RBC cholinesterase is protective against a 10% change in brain cholinesterase activity by a factor of at least 4.

71.7.6.1  Carcinogenicity The liver and nasal tumors observed in the rat and mice studies have been submitted to independent peer reviews and discussions by regulatory authorities. Because regulatory agencies interpret the same set of data according to slightly different criteria, the outcome of such reviews can vary in different areas of the world. The U.S. EPA has classified malathion in the category “Suggestive Evidence for Carcinogenicity” (i.e., not sufficient evidence to definitely conclude whether malathion can be carcinogenic or not). In the European Union, “no classification with regards to carcinogenicity” has been proposed by the experts (EFSA, 2004, 2006; UK COM/COC, 2003; U.S. EPA, 2006b). In conclusion, malathion may cause an increase in benign but not malignant tumors at dose levels high enough to produce organ toxicity, an effect that occurs via a nongenotoxic mechanism, and for which a clear NOAEL can be identified. These tumors occurred at dose levels higher

than the NOAEL/BMDL20 established for RBC cholinesterase activity. Therefore, risk assessments conducted using a NOAEL/BMDL20 for RBC cholinesterase activity will be protective of any tumorigenic effect.

71.7.7  Reproductive and Developmental Effects Malathion has been tested for reproduction toxicity in the rat and for developmental toxicity and teratogenicity in the rat and rabbit. Malathion demonstrated no teratogenic potential in either species, and it was not toxic to reproduction. Malathion may result in weight reductions in offspring when dams receive high dose levels. The potential for reproductive toxicity was investigated in a two-generation reproduction study in rats with four dose levels from 550 to 7500 ppm. The NOAEL for reproductive and parental toxicity was 7500 ppm (595 mg/ kg BW/day), the highest dose tested. Based on slightly reduced body weights in offspring from dams dosed at 5000 ppm, the offspring NOAEL was concluded to be 1700 ppm (132 mg/kg/day). In a rat prenatal developmental toxicity study with dose levels of 200, 400, and 800 mg/kg BW/day, the NOAEL for maternal toxicity was found to be 400 mg/kg BW/day based on slightly reduced body weight gains and urinestained fur at higher dose levels. There were no treatmentrelated fetal effects. Thus, the NOAEL for developmental effects was 800 mg/kg BW/day, the highest dose tested. A rabbit developmental toxicity study was also conducted, and the NOAEL for maternal toxicity was found to be 25 mg/kg BW/day based on decreased body weight gain at the higher dose levels. There was a slightly increased incidence in mean number and percentage of resorptions at the maternotoxic dose levels of 50 and 100 mg/kg BW/ day. There was no difference in fertility, number of corpora lutea, implantation sites, litter size, or fetal weight and length. No other sign of toxicity was seen in dams or fetuses, nor was there any evidence for teratogenicity. When the data from the main study and range finder studies were combined, there was no effect on postimplantation

Table 71.5  Meta-analysis for RBC and Brain Cholinesterase Data in Rats After Chronic (Lifetime) Exposure to Malathion Sex

Brain ChEI

RBC ChEI

BMD10 (mg/kg/day)

BMDL10 (mg/kg/day)

BMD20 (mg/kg/day)

BMDL20 (mg/kg/day)

Male

231.1

128.1

96.2

52.4

Female

337.7

233.0

104.0

59.3

Chapter | 71  Malathion: A Review of Toxicology

loss, and the live litter size was higher in all treated groups compared to control. Therefore, it can be concluded that the NOAEL for fetal toxicity is 100 mg/kg BW/day, the highest dose tested (Joint FAO/WHO Meeting on Pesticide Residues, 2003). However, the increased incidence in resorption may also be interpreted as an adverse effect of the treatment, and the NOAEL will then be lower – that is, 25 mg/kg BW/day (EFSA, 2006; U.S. EPA 2006a).

71.7.8  Neurotoxicity A full complement of neurotoxicity studies have been conducted on malathion, including an acute delayed neurotoxicity study in hens, acute and subchronic toxicity studies in rats, and a developmental neurotoxicity study in rats. In most studies, malathion produced only mild to moderate signs of cholinergic toxicity. Malathion does not cause permanent nerve damage and did not induce acute delayed neurotoxicity. The developmental neurotoxicity (DNT) study demonstrated that malathion is not a developmental neurotoxicant. The ability to cause acute delayed neurotoxicity was investigated in a study with the hen. Delayed neurotoxicity is a serious, sometimes irreversible, condition in which nerve damage may occur after acute or chronic exposure to certain toxicants leading to sensory loss, progressing muscle weakness of the extremities, and ataxia. The potential to induce delayed neurotoxicity depends on the ability of a chemical to interfere irreversibly with specific enzymes, namely neurotoxicity target esterases. Due to the seriousness of the condition, all organophosphate-based plant protection products must be specifically tested for delayed neurotoxicity before commercialization. The hen was chosen by regulators as the test animal because it is particularly sensitive to this effect. Birds are treated with a very high dose and thereafter examined for specific signs of nerve damage. To get the dose as high as possible, hens are simultaneously given an organophosphate antidote (atropine). This makes treatment at dose levels above the LD50 possible. In the delayed neurotoxicity study with malathion, birds received a single dose of 1.3 times the LD50 (~1000 mg/ kg BW/day). Three weeks after the first dose, birds were dosed again, this time at 1.1 times the LD50 (~850 mg/kg BW/day). To ensure survival, they were treated with atropine sulfate before and after dosing. Thirty-eight out of 60 birds died within 2 weeks after the first dosing, and 7 died after the second dose. The remaining birds survived until scheduled sacrifice on day 42. Signs of toxicity included lethargy, general weakness, ataxia, anorexia, inability to stand, diarrhea, sitting on haunches, paralysis of legs and wings, and pale comb. Survivors recovered completely within 6 days after dosing. Positive control birds showed clinical signs of delayed neurotoxicity as well as changes typical of delayed polyneuropathy in the spinal cord and

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the sciatic nerve. No clinical signs of delayed neurotoxicity and no histopathological neural changes were found in any of the malathion-treated birds. Based on the results of the previously discussed studies, it has been concluded that malathion is not an inducer of delayed neurotoxicity. In the acute neurotoxicity study in rats, adult rats were given single oral doses from 500 to 2000 mg/kg malathion in corn oil. Functional observational battery (FOB), locomotor activity, histopathology, and cholinesterase inhibition assays were performed at pretest, peak effect (15 min postdosing), day 7, and day 14. Treatment-related effects on behavioral parameters were minimal at even the highest dose tested. Nerve fiber degeneration was noticed slightly more frequently in the 2000 mg/kg BW males than in the control males. In all cases, the findings were observed in single animals and graded as minimal. Furthermore, it was within the range of expected occurrence based on a comparison with historical control data. Therefore, it was concluded to be spontaneous and not related to treatment. Plasma and RBC cholinesterase inhibition results were highly variable. Statistically significant inhibition of RBC cholinesterase inhibition was seen at the highest dose. Brain cholinesterase was not inhibited at any dose level. In the subchronic (13 weeks) dietary neurotoxicity study with malathion in rats, the highest dose level was 20,000 ppm (~1500 mg/kg BW/day). The only treatmentrelated clinical signs noted were yellow or orange material in the anogenital or urogenital regions and on the tail in high-dose animals. No treatment-related neural lesions were found in rats receiving 20,000 ppm malathion/day in the diet. The NOAEL for brain AChE inhibition was 5000 ppm (~350 mg/kg BW/day). Based on inhibition of RBC cholinesterase at 5000 ppm, the NOAEL from this study was 50 ppm (4 mg/kg BW/day for males and females).

71.7.8.1  Developmental Neurotoxicity and Relative Sensitivity Two studies – a DNT study and a relative sensitivity cholinesterase inhibition study – were conducted to investigate whether fetuses and juvenile animals are more sensitive to malathion exposure compared to dams or other adult animals. (a) Developmental Neurotoxicity The DNT study was designed to assess the potential of malathion to cause functional or morphological changes in the developing nervous system of rat offspring following maternal exposure during pregnancy and early lactation and following direct exposure of the offspring from early lactation and until weaning. Malathion was given by oral gavage in corn oil at dose levels ranging from 5 to 150 mg/kg BW/day. In dams, daily consecutive doses were given from gestation day (GD) 6 to day 10 of lactation. Hereafter, dosing of the dams stopped

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and instead the offspring received 11 consecutive doses from postnatal day (PND) 11 to PND 21. From birth until study termination on PND 65, offspring were submitted to various investigations at selected time points, including age-appropriate FOBs, motor activity assessments, learning and memory tests, brain histopathology and morphometrics, and assessments of sexual maturation. FOB assessments were also performed on the dams during gestation and lactation. Adverse effects occurred only in the high-dose group during the period of direct repeat-dose treatment and only in four offspring, all from the same litter. The effects noted in these animals included postdosing clinical signs such as whole-body tremors, underactivity, prostate posture, partially closed eyelids, and abnormal gait. These effects were considered to be a direct effect of dosing and not a developmental effect. On PND 11, a female offspring in the highest dose group failed to show an immediate surface righting reflex. This was not seen in any of the other arena observations before or after PND 11. Overall, the DNT study demonstrated no evidence of any permanent functional changes in the development of the nervous system. Based on slower righting effect and postdosing clinical signs in offspring at 150 mg/kg BW/day, the NOAEL was 50 mg malathion/kg BW/day for morphological and functional development of the nervous system. The NOAEL for maternal toxicity was 150 mg/kg BW/day. (b) Relative Sensitivity/Cholinesterase Study The relative sensitivity (cholinesterase inhibition) study was conducted to investigate whether fetuses, nursing offspring, and juvenile animals are more sensitive to malathion than dams or other adult animals. In this study, groups of animals received malathion by oral gavage in corn oil at repeated doses ranging from 5 to 150 mg/kg BW/day or as single doses ranging from 5 to 450 mg/kg BW. Dams were given daily consecutive doses from GD 6 to GD 20 or from GD 6 to PND 10. One subset of pups was given a single dose on PND 11 (these pups were selected from nondosed dams). Another subset of pups were given 11 consecutive daily doses from PND 11 to PND 21 (these pups were selected from dams given the same dose; i.e., they had already been treated in utero and during lactation). The young adult rats (naïve animals, 7 or 8 weeks old) were given either 1 single dose or 11 consecutive doses. The rats were dosed at approximately the same time every day. Cholinesterase activity was measured in gestating females and their preterm fetuses on GD 20; in preweaning offspring on PND 4, 11, and 21; and on PND 60 in young adult rats. Inhibition of RBC cholinesterase activity in dams receiving 150 mg/kg BW/day was markedly greater than

Hayes’ Handbook of Pesticide Toxicology

that for their fetuses at GD 20. At lower dose levels, the magnitude of the effects on blood cholinesterase levels was generally similar in dams and fetuses, with inhibition occurring at 50 mg/kg BW/day. Cholinesterase activity for PND 4 offspring was unaffected up to and including maternal doses of 150 mg/kg BW/day. For offspring at PND 11, after a single dose, blood cholinesterase was inhibited at and above 50 mg/kg BW/day, whereas brain cholinesterase was inhibited at 150 mg/kg BW/day. The reaction of the young adults was less marked with a NOAEL of 150 mg/kg BW/day for RBC cholinesterase and 450 mg/kg BW/day for brain cholinesterase. The overall NOAEL for RBC cholinesterase inhibition was 5 mg/kg BW/day, and that for brain cholinesterase was 50 mg/kg BW /day. The inhibition of RBC and brain cholinesterase activity was compared for each subpopulation using by the BMD approach. Key results from these comparisons are summarized in Table 71.6. Overall, it can be concluded (1) that RBC cholinesterase inhibition was the most sensitive compartment for all subpopulations with the exception of the fetuses, in which the responses in the RBC and brain compartment appear to be similar based on a comparison of BMD10 values; (2) that there was no consistent pattern of sensitivity observed in males and females; (3) that the fetuses were equally or less sensitive than the dams; (4) that the cholinergic response in nursing (PND 4) offspring was much lower than that in the dams, indicating that nursing offspring receive little or no malathion via the milk produced by the dams; and (5) that very young (PND 11) rat offspring were approximately 10 times more sensitive to direct malathion exposure compared to adults, but this sensitivity decreased considerably to approximately 2 times that of adults as the offspring matured during the next 10 days.

71.7.9  Human Volunteer Study A human volunteer study has been conducted. The study was designed as a randomized, double-blind study and followed the procedures routinely used for evaluating drug safety (conducted in compliance with the Declaration of Helsinki and with the principles of Good Clinical Practice). Fully informed volunteers received single doses of malathion in gelatin capsules containing dose levels from 0.5 to 15 mg/kg BW or placebo. An extensive range of clinical observations were performed at screening, predose, and 24 h after dosing, and adverse clinical events were recorded up to 2 weeks after dosing and fully described with respect to their duration and severity. An assessment was made as to their possible treatment relatedness and details of any specific therapy required were recorded.

Chapter | 71  Malathion: A Review of Toxicology

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Table 71.6  Benchmark Doses Calculated by the U.S. EPA for RBC and Brain Cholinesterase Data Obtained from a Relative Sensitivity (Comparative Cholinesterase) Study Subpopulation

RBC Cholinesterase

Brain Cholinesterase

Male

Female

Male

Female

Offspring, single oral gavage dose on PND 11 (mg/kg BW)

BMD20  41.3 BMD10  16.9

BMD20  33.7 BMD10  18.1

BMD10  24.6

BMD10  23.6

Adult, single oral gavage dose (mg/kg BW)

BMD20   150a BMD10  158

BMD20  150a BMD10  150a

BMD10  150a

No inhibition observed, so BMDs not calculated

Offspring, 11 daily oral BMD20  21.5 gavage doses from PND 11 BMD10  12.7 to PND 21 (mg/kg/day)

BMD20  28.1 BMD10  13.6

BMD10  85.7

BMD10  91.2

Adult, 11 daily oral gavage BMD20  53 doses (mg/kg/day) BMD10  23.0

BMD20  47 BMD10  22.7

BMD10  150a

BMD10  150a

Dams, daily oral gavage dosing from GD 6 to GD20 (mg/kg/day)

BMD20  41 BMD10  21.1

NA

BMD10  150a

NA

Fetuses, daily oral gavage BMD20  150a dosing to the dam GD 6 to BMD10  77.8 GD20 (mg/kg/day) Offspring, PND 4 (mg/kg/ day)

BMD10  150a

Offspring, 11 daily oral BMD20  28.1 gavage doses from PND 11 BMD10  12.7 to PND 21 (mg/kg/day)

BMD10  70.4

BMD10  150a

BMD10  150a

BMD10  150a

BMD20  30.3 BMD10  13.6

BMD10  85.7

BMD10  91.2

NA, not applicable. a Calculated value is outside of tested dose range.

Blood cholinesterase activity was measured; the study design had sufficient statistical power to detect a dose that would cause cholinesterase inhibition in excess of 15%. Blood samples were analyzed for malathion and malaoxon on many sampling occasions, and urine samples were also collected for determination of malathion metabolites. Malathion was tolerated well from 0.5 to 15 mg/kg BW in male subjects and at 15 mg/kg BW in female subjects. No clinically significant changes in vital signs, ECGs, hematology, clinical chemistry, urinalysis, and physical parameters were observed in any subjects during the study. There were no adverse effects or other effects of malathion in any subjects and no statistically significant reductions in plasma or RBC cholinesterase activity in male or female subjects. A number of adverse events were recorded; they were all mild and the frequency of occurrence (35–40%) was similar in volunteers receiving placebo or malathion. Headache was the most commonly reported adverse event,

occurring intermittently in 20–30% of the volunteers during the 2-week follow-up period. None of these adverse events were considered related to the administration of malathion. At the highest dose level of 15 mg/kg BW, plasma levels of malathion and malaoxon remained below detectable levels (less than ~100 ng/ml plasma) at all times after administration, indicating the degree of protection offered by first-pass hepatic metabolism. Approximately 90% of malathion metabolites were excreted in urine within 12 h of dosing, and excretion was essentially complete within 24–48 h of dosing. The major metabolites in urine were MMCA and MDCA (accounting for up to approximately 30 and 13%, respectively, of the administered dose). Dimethyl thiophosphate, dimethyl phosphate, and dimethyl dethiophosphate were also measured in the urine and together accounted for approximately 30% of the total administered dose. In conclusion, based on no effects on plasma and RBC cholinesterase and no clinical effects of treatment, the highest dose tested of 15 mg/kg BW was a clear NOAEL

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in both sexes. This dose results in nondetectable plasma levels of malathion and malaoxon.

71.7.10  Toxicology Studies on Malaoxon Malaoxon is formed by oxidation of malathion in mammals and therefore its toxicity is inherently tested as part of any study on malathion. Nevertheless, because malaoxon is also formed by oxidation of malathion in the environment, and because it is the only metabolite of malathion that is more toxic than malathion, several studies have been conducted to describe the acute, short-term and long-term toxicity of malaoxon.

71.7.10.1  Acute Toxicity The relative potency of malaoxon compared to malathion was investigated in an acute dose study using 11-day-old rat offspring. Rats were given a single oral gavage dose of malathion (at dose levels ranging from 10 to 150 mg/kg BW) or malaoxon (from 10 to 12.5 mg/kg BW). RBC and brain cholinesterase activities were measured at the time of peak effect, which was determined for each compound in preliminary studies. There were no mortalities. Mild clinical signs occurred in malaoxon-treated pups at and above 10 mg/kg BW and in malathion-treated pups at and above 100 mg/kg BW. The RBC and brain cholinesterase data from this study were analyzed using BMD methodology to estimate the relative potency of malaoxon compared to malathion. Relative potency factors for malaoxon in each cholinesterase compartment were calculated for both males and females by dividing the BMDs for malathion by the BMDs for malaoxon. A summary of the BMDs and potency factors is provided in Table 71.7. A comparison of the BMDs calculated for cholinesterase inhibition in this study for malaoxon and malathion established an acute relative potency factor of 223 for malaoxon.

71.7.10.2  Short-Term Exposure The potency of malaoxon to inhibit cholinesterase activity after repeated dosing was compared to that for malathion in a study using 11-day-old rat offspring. In this study, malathion and malaoxon were administered to rat offspring by oral gavage daily from PND 11 through PND 21. RBC and brain cholinesterase activities were measured on PND 21 at the time of peak effect. Dose levels ranged from 5 to 150 mg/kg BW for malathion and from 0.1 to 4 mg/kg BW for malaoxon. There were no mortalities and no effects on body weight or body weight gain in any group. Treatment with malathion at 150 mg/kg/day caused adverse clinical signs consistent with cholinergic poisoning. RBC cholinesterase levels were reduced after treatment with malathion at and above 25 mg/kg/day, and they were reduced at and above 1 mg/kg/day in the malaoxon-treated groups. Brain cholinesterase levels were reduced only in the 150 mg/kg/day dose group administered malathion. There was no inhibition of brain cholinesterase levels for pups administered malaoxon. The RBC and brain cholinesterase data from this study were analyzed using the U.S. EPA’s BMD methodology to estimate the relative potency of malaoxon compared to malathion. Relative potency factors for malaoxon in each cholinesterase compartment were calculated for both males and females by dividing the BMDs for malathion by the BMDs for malaoxon. A summary of the BMDs and potency factors is provided in Table 71.8. A comparison of the BMDs calculated for cholinesterase inhibition in this study for malaoxon and malathion established a relative potency factor of 333 for repeated doses of malaoxon. The toxicity of malaoxon has also been evaluated in a 14-day dietary toxicity study in adult rats. In this study, the NOAEL was 25 ppm (~3 mg/kg BW/day). RBC cholinesterase inhibition occurred at dose levels of 100 ppm and higher, and brain ChE was inhibited at 2500 ppm and higher. The highest dose employed was 3500 ppm (295 mg/ kg BW/day). At this dose, malaoxon produced excessive

Table 71.7  Comparison of the Relative Potency of Malaoxon and Malathion to Inhibit Cholinesterase Activity After a Single Oral Gavage Dose to Rat Offspring Sex

Compartment

Malathion BMD10 (mg/kg BW)

Malaoxon BMD10 (mg/kg BW)

Relative Potency Factor of Malaoxona

Male

RBC Brain

10.8 41.6

0.50 2.8

21.6 14.9

Female

RBC Brain

12.3 39.6

0.69 3.6

17.8 11.0

a

The relative potency factor was calculated using the malathion most sensitive endpoint of RBC ChEI in males (10.8) divided by the malaoxon endpoint (0.50); therefore, 10.8/0.50 ≈ 21.6 ≈ 22.

Chapter | 71  Malathion: A Review of Toxicology

toxicity, as evidenced by clinical signs consistent with cholinesterase poisoning and mortality.

71.7.10.3  Long-Term Toxicity In a 24-month chronic toxicity/carcinogenicity study, rats received malaoxon at dietary dose levels of 0, 20, 1000, and 2000 ppm. Treatment at 1000 and 2000 ppm had an effect on survivorship. The most common causes of death were purulent inflammation of the lungs and mononuclear cell leukemia, one of the most common neoplasms in this particular strain of rats. In males, the incidence of mononuclear cell leukemia followed a dose-related trend; however, it was still within the historical control levels of the testing laboratory. No statistically significant increase was evident in pairwise comparisons between treated groups and controls. Food consumption, body weight, and body weight gain were affected at mid- and high doses. RBC and brain ChE activities were inhibited at 1000 and 2000 ppm; RBC ChE was inhibited to a higher degree than brain ChE. In addition, inflammatory changes were seen in nasal turbinate, lungs, and tympanic membrane at the high doses. Inhaled food particles were found to be retained in the nasal cavity, and the observed inflammation was considered to be secondary to the food particle deposition and response to injury due to the presence of foreign material. The NOAEL was 20 ppm, approximately 1 mg/kg BW/day. BMD modeling has been performed using the RBC and brain ChE data from this study. A summary of the BMD analysis is provided in Table 71.9. The modeling demonstrated that RBC ChE inhibition was much more sensitive compared to brain ChE inhibition, and that males may be slightly more sensitive to malaoxon exposure compared to females. The lowest BMDL20 value was 0.5 mg/kg/day for male RBC ChE inhibition; this value may be considered the equivalent to a study NOAEL. The U.S. EPA also calculated BMD10s for RBC ChE inhibition in this study on malaoxon and compared it to BMD10

1541

values calculated for RBC ChE inhibition from data obtained from a similar chronic toxicity/oncogenicity study on malathion in order to establish a chronic relative potency factor for malaoxon (U.S. EPA, 2006a). This comparison revealed that malaoxon is 61 times more potent as a ChE inhibitor compared to malathion on a chronic (lifetime) exposure basis. The U.S. EPA has fully accounted for the increased chronic potency of malaoxon in its risk assessments.

Conclusions Malathion has been used as an insecticide with a variety of applications for approximately 50 years. During this time, its toxicology has been extensively studied and reported. Because malathion exhibits low mammalian toxicity and it has been shown that relevant impurities may enhance toxicity, it is essential that the impurity profile of malathion used in toxicity studies is known. Malathion is a pro-insecticide, and the necessary bioactivation to malaoxon results in a delay in which other metabolic reactions take place. Consequently, the formation of MMCA and MDCA dominates its metabolism in mammals. The most sensitive effect of malathion is the inhibition of AChE activity, with RBC AChE being more sensitive than brain AChE. This effect was observed in rat, mouse, dog, and human. Malathion does not cause permanent nerve damage. Clinical signs of toxicity are mild and generally occur at higher dose levels. Effects at low dose levels are minimal or not observed. Malathion is not genotoxic in vivo, and it does not have an adverse impact on reproduction. Slightly reduced pup weights were seen at high dose levels. Because of the sensitivity of RBC AChE to malathion exposure, most regulatory bodies have selected it as the endpoint for use in risk assessment, to be protective of other endpoints of concern. Endpoints for risk assessments may be based on NOAELs or may be based on calculated BMDs (e.g., 20% inhibition of RBC AChE).

Table 71.8  Comparison of the Relative Potency of Malaoxon and Malathion to Inhibit Cholinesterase Activity After 11 Days of Daily Oral Gavage Doses to Rat Offspring Sex

Compartment

Malathion BMD10 (mg/kg/day)

Malaoxon BMD10 (mg/kg/day)

Relative Potency Factor of Malaoxona

Male

RBC Brain

12.84 96.67

0.3929 NAb

32.7 —

Female

RBC Brain

13.12 76.84

0.4982 NAb

26.3 —

NA, not available. a The relative potency factor was calculated using the malathion most sensitive endpoint of RBC ChEI in males (12.84) divided by the malaoxon endpoint (0.3929); therefore, 12.84/0.3928 ≈ 32.7 ≈ 33. b BMDs for brain cholinesterase were not calculated because no inhibition was observed at the doses tested.

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Table 71.9  Benchmark Dose Analysis for Red Blood Cell and Brain Cholinesterase Data in Rats After Chronic (Lifetime) Exposure to Malaoxon Sex

Brain ChEI

RBC ChEI

BMD10 (mg/kg/day)

BMDL10 (mg/kg/day)

BMD20 (mg/kg/day)

BMDL20 (mg/kg/day)

Male

96.8

87.8

3.3

0.5

Female

92.1

86.4

0.9

0.5

Acknowledgments The authors thank Dr. Richard Reiss (Exponent, Inc.) for his contribution on benchmark dose analysis, Dr. Judith Hauswirth for input on toxicology data, and Dr. Terry Roberts for his advice and editorial assistance.

References Aldridge, W. N., Miles, J. W., Mount, D. L., and Verschoyle, R. D. (1979). The toxicological properties of impurities of Malathion. Arch. Toxicol. 42, 95–106. Ambrus, A., Hamilton, D. J., Kuiper, H. A., and Racke, K. D. (2003). Significance of impurities in the safety evaluation of crop protection products. Pure Appl. Chem. 75, 937–973. Baker, E., Warren, M., Zack, M., Dobbin, R., Miles, J., Miller, S., Alderman, L., and Teeters, W. (1978). Epidemic malathion poisoning in Pakistan malaria workers. Lancet 1(8054), 31–34. European Food Safety Authority (EFSA) (2006). Conclusion regarding the peer review of the pesticide risk assessment of the active substance malathion. Finalized: 13 January 2006, No. 63, 1–87. Available at http://www.efsa.europa.eu. European Food Safety Authority (EFSA) (2004). Draft Assessment Report: Public version: Initial risk assessment provided by the rapporteur Member State Finland for the existing active substance Malathion of the second stage of the review programme referred to in Article 8(2) of Council Directive 91/414/EEC, Vol. 3, Annex B, part 2, B.6. Available at http://dar.efsa.europa.eu/dar-web/provision. Joint FAO/WHO Meeting on Pesticide Residues (1997). Pesticide Residues in Food—1997 Evaluations Part II. Toxicological and Environmental, WHO/PCS/98.6, 1998, Nos. 924-942 on INCHEM. World Health Organization, Geneva. Available at http://www.who. int/ipcs/publications/jmpr/en. Joint FAO/WHO Meeting on Pesticide Residues (1999). Pesticide Residues in Food—1999 Evaluations Part I. Residues, FAO Plant Production

and Protection paper 157, 2000. World Health Organization Geneva. Available at http://www.who.int/ipcs/publications/jmpr/en. Joint FAO/WHO Meeting on Pesticide Residues (2003). Pesticide Residues in Food—2003 Evaluations Part II. Toxicological evaluations, WHO/ PCS/04.1. World Health Organization, Geneva. Available at http:// www.fao.org/ag/agp/agpp/pesticid/jmpr/DOWNLOAD/2003.pdf. Roberts, T. R., and Hutson, D. H. (1999). “Metabolic Pathways of Agro­ chemicals, Part 2, 364.” Royal Society of Chemistry, Cambridge, UK. UK COM/COC (2003). Joint statement on review of malathion: COM/03/S1 & COC/03/S1—March 2003. Available at http://www. iacoc.org.uk/statements/jointstatementonreviewofmalathioncom03s1andcoc03s1.htm. Umetsu, N., Grose, F. H., Allahyari, R., Abu-El-Haj, S., and Fukuto, T. R. (1977). Effect of impurities on the mammalian toxicity of technical malathion and acephate. J. Agric. Food Chem. 25(4), 946–953. U.S. Environmental Protection Agency (2000). Background information about the benchmark dose model used by EPA, as well as instructions for downloading and using the model http://www.epa.gov/NCEA/ bmds/index.html. Additional guidance from EPA can be found at: http://www.epa.gov/ncea/pdfs/bmds/BMD-External_10_13_2000.pdf. U.S. Environmental Protection Agency (2006a). United States Environmental Protection Agency, Office of Prevention, Pesticides and Toxic Substances, Memorandum dated June 31, 2006: Malathion: Revised Human Health Risk Assessment for the Reregistration Eligibility Decision Document (RED), PC Code 057701, Case No. 0248, DP Barcode D330680. Available at http://www.regulations. gov/fdmspublic/component/main?main  DocketDetail&d  EPAHQ-OPP-2004-0348. U.S. Environmental Protection Agency (2006b) United States Environmental Protection Agency, Office of Prevention, Pesticides and Toxic Substances, Memorandum dated July 31, 2006 Malathion: Revised Human Health Risk Assessment for the Reregistration Eligibility Decision Document (RED), PC Code 057701, Case No. 0248, DP Barcode D330680. Available at http://www.epa. gov/oppsrrd1/REDs/malathion_red. World Health Organization (1999). “JMPR report. Interpretation of cholinesterase inhibition data.” World Health Organization, Geneva.

Chapter 72

Clinical Toxicology of Anticholinesterase Agents in Humans Marcello Lotti Università degli Studi di Padova

Several chemicals display anticholinesterase activity among which organophosphorus esters (OPs) represent the vast majority because they are widely used and easily available. Since their introduction after World War II as insecticides (Khurana and Prabhakar, 2000), countless publications have described their clinical toxicology in humans. Initially, most case reports dealt with accidental and occupational poisoning, whereas later the majority of poisonings were the result of suicide attempts. Most recently, attention has been directed to possible subtle effects caused by low-level long-term exposures such as those encountered in the workplace. Concurrently, an even larger amount of experimental studies have contributed to our understanding of anticholinesterase toxicology. Therefore, anticholinesterase agents, and OPs in particular, undoubtedly represent the most extensively studied class of chemicals in toxicology and it is not surprising that a very large number of reviews, textbooks, book chapters, and other general publications, in addition to research papers, have appeared over the years on their clinical toxicology (the most recent reviews include: Bardin et al., 1994; Brown and Brix, 1998; De Bleecker et al., 1992a; ECETOC, 1998; Eyer, 1995; Gunderson et al., 1992; Jamal, 1997; Karalliede, 1999; Karalliedde and Senanayake, 1999; Marrs, 1993; Millard and Broomfield, 1995; Ray, 1998a, b; Steenland, 1996). Nevertheless, some controversies still exist on some aspects of both the clinical toxicology and the treatment of OP poisoning. In particular, conflicting results have been published on long-term sequelae of acute exposure and on possible effects of longterm exposures that do not cause overt cholinergic toxicity, and different opinions exist on the uses of antidotes in the treatment of acute poisonings. This chapter deals with the clinical aspects of OP toxicology in humans, mentioning, where appropriate, those of other anticholinesterase agents and supporting experimental evidence in animals as well. Moreover, this chapter Hayes’ Handbook of Pesticide Toxicology Copyright © 2001 Elsevier Inc. All rights reserved

does not replace the specific chapters of the previous editions of this book (Hayes, 1982; Hayes and Laws, 1991), where additional details on OP toxicology, as derived from the older literature, can be found.

72.1  The cholinergic syndrome The cholinergic syndrome is characterized by overstimulation of cholinergic receptors throughout the body. It may be the consequence of single or repeated exposures to a variety of chemicals such as OPs and carbamates, which inhibit acetyl-cholinesterase (AChE) at the synaptic level. As a consequence, the level of acetylcholine increases, although the amount and time course of neurotransmitter accumulation may vary widely among various districts of the cholinergic system (Stavinoha et al., 1976). Thus, the clinical picture that results from this excess of neurotransmitter may be quite variable in presentation of signs and symptoms, at the onset, in time course, and in outcome. This clinical polymorphism largely depends on the chemical involved and on the dose.

72.1.1  Etiology 72.1.1.1  Anticholinesterases as Pesticides and Warfare Agents Although the use of anticholinesterase OP insecticides has declined during the last two decades, particularly in agriculture, they still represent an important class of pesticides, which account for about 10% of all active ingredients currently used as pesticides. However, several OPs that caused poisoning in humans in the past are believed to be no longer manufactured or marketed (Tomlin, 1997). The long inventory of acute intoxications in humans involves a variety of OPs. Table 72.1 lists examples of OPs, 1543

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Table 72.1  Classification of Organophosphorus Esters;a General Formula: R1 R2

O P

X

A. Compounds where X  halogen or CN, CNS, etc. 1. R1  alkoxy, R2  alkyl Example: sarin (isopropyl methylphosphonofluoridate), nerve gas (Okumura et al., 1996) 2. R1 and R2  alkoxy Example: DFP (diisopropyl phosphorofiuoridate), laboratory chemical and discontinued drug (Moore, 1956) 3. R1  alkylamide, R2  alkoxy Example: tabun (ethyl N-dimethylphosphoroamido cyanidate), nerve gas (Compton, 1987) 4. R1 and R2  mono- or dialkylamido Example: mipafox (N N-diisopropyl phosphorodiamido fluoridate), laboratory chemical (Bidstrup et al., 1953). B. Compounds where X  alkyl, alkoxy, or aryloxy 1. Alkoxydialkyl or dialkoxyalkyl compounds Example: trichlorfon (dimethyl (2,2,2-trichloro-l-hydroxyethyl) phosphonate), insecticide and drug (Vasilescu and Florescu, 1980) 2. Trialkyl compounds and dialkoxy, aryloxy compounds Example: dichlorvos (dimethyl 2,2 dichlorovinyl phosphate), insecticide (Wadia et al., 1987) C. Thiol- and thiono-phosphorus compounds 1. Thiol compounds Example: demeton-S-methyl (S-[2-(ethylthio)ethyl] dimethyl phosphorothioate), insecticide (Weir et al., 1992) 2. Thiono compounds Example: parathion (diethyl O-(4-nitrophenyl)phosphorothioate), insecticide (Namba et al., 1971) 3. Thiol – thiono compounds Example: malathion (dimethyl S-(l,2-dicarboxylethyl) phosphorodithioate), insecticide (Baker et al., 1978) D. Derivatives of pyrophosphorous acid Example: sulfotepp (tetraethyl dithionopyrophosphate), insecticide (Namba et al., 1971) E. Compounds containing a quaternary nitrogen Example: ecothiophate (diethyl-S-2-trimethyl ammonium-ethyl phosphorothioate iodide), drug (Gesztes, 1966). a

Nerve gases belong to group A, whereas most common pesticides belong to groups B and C. Examples of OPs which caused poisoning in humans are given. Source: Holmstedt (1963).

ranked according to their chemical structure, that caused acute toxicity in humans. Most compounds that belong to group A are chemical warfare agents and are highly toxic. Most pesticides belong to groups B and C, among which dimethoxy OPs are the most common. In contrast to other OPs, dimethoxy OPs are less toxic because of certain biochemical characteristics (i.e., relative high speed of reactivation of dimethoxyphosphorylated AChE; see Section 72.1.2.2). In Table 72.2, the World Health Organization (WHO) recommended classification of OP pesticides by hazard is reported. The classification is based on acute oral and dermal toxicity in rats and the physical state of product or formulation. Confirmation of the hazard severity of these chemicals has often been obtained from clinical observations. OPs have also been used as warfare agents and large amounts are thought to be stockpiled in arsenals worldwide. They were developed during the 1950s and were used as recently as the early 1990s in a terrorist attack in a Tokyo subway (Suzuki et al., 1995). Moreover, during the Persian Gulf War, soldiers were thought to have been exposed to low levels of nerve gases (NIH, 1994).

72.1.1.2  Anticholinesterases as Drugs Anticholinesterase drugs have a variety of indications in clinical medicine, and toxicity may arise from their therapeutic uses. However, although used in the past, nowadays drugs do not include OPs, but use reversible inhibitors of AChE such as physostigmine, pyridostigmine, neostigmine, and edrophonium. Their acceptability has been established in four areas: atony of the smooth muscle of the intestinal tract and urinary bladder, myasthenia gravis, glaucoma, and termination of effects of competitive neuromuscular blocking drugs (Taylor, 1996a). Moreover, reversible anticholinesterases are also used for treatment of overdoses of atropine and other anticholinergic drugs (such as phenothiazines and tricyclic antidepressant; Nilsson, 1982). Pyridostigmine was administered on a large scale during the Persian Gulf War to soldiers as a prophylaxis for nerve gas attacks to obtain reversible carbamylation of AChE, thereby preventing irreversible phosphorylation (Keeler et al., 1991). Several AChE inhibitors have been used or are on clinical trial for the treatment of Alzheimer’s disease. The rationale

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Table 72.2  Organophosphorus Pesticides and Risk of Cholinergic Syndromea Classification

Common name

Extremely hazardous

Chlorethoxyfos, chlormephos, coumaphos, disulfoton, EPN, ethoprophos, fenamiphos, fonofos, mevinphos, parathion, parathion-methyl, phorate, phosphamidon, sulfotepp, tebupirimfos, terbufos

Highly hazardous

Azinphos-ethyl, azinphos-methyl, cadusafos, chlorfenvinphos, demeton-S-methyl, dichlorvos, dicrotophos, edifenphos, famphur, heptenophos, isazofos, isofenphos, isoxathion, mecarbam, methamidophos, methidation, monocrotophos, omethoate, oxydemeton-methyl, pirimiphos-ethyl, propaphos, propetamphos, thiometon, triazophos, vamidothion

Moderately hazardous

Anilofos, bilanafos, butamifos, chlorpyrifos, cyanophos, diazinon, dimethoate, ethion, etrimfos, fenitrothion, fenthion, formothion, methacrifos, naled, phenthoate, phosalone, phosmet, phoxim, piperophos, profenofos, prothiofos, pyraclofos, pyrazophos, quinalphos, sulprofos

Slightly hazardous

Acephate, azamethiphos, iprobenfos, malathion, pirimiphos-methyl, pyridaphenthion, trichlorfon

a

WHO recommended classification by hazard (WHO, 1998). Based on acute oral and dermal toxicity in rats and the physical state of product or formulation, pesticides are ranked as follows: Extremely hazardous: 5 mg/kg or less if solid; 20 mg/kg or less if liquid Highly hazardous: 5–50 mg/kg if solid; 20–200 mg/kg if liquid Moderately hazardous: 50–500 mg/kg is solid; 200–2000 mg/kg if liquid Slightly hazardous: over 500 mg/kg if solid; over 2000 mg/kg if liquid Only compounds in use or being developed are reported (Tomlin, 1997).

increasing synaptic levels of acetylcholine (Summers et al., 1986). Tacrine (tetrahydroaminoacridine) is the most extensively studied drug; however, controversies arose from initial reports (Relman, 1991) and clinical results have so far not been convincing (Davis et al., 1992; Maltby et al., 1994). Finally, pyridostigmine salicylate was shown to rapidly reverse clinical signs of central and peripheral anticholinergic toxicity caused by a variety of drags, such as antidepressants, antiparkinsonians, antihistamines, and antispasmodics, and toxic plants, such as mushrooms, potato sprouts, and bittersweet (Granacher and Baldessarini, 1975).

72.1.2  Pathogenesis OPs exert their toxic action by interfering with cholinergic transmission. The molecular mechanism of cholinergic toxicity involves the interaction of OPs with AChE, which almost completely explains all the signs and symptoms of acute OP poisoning. Symptoms and signs are related to excess acetylcholine and the consequent overstimulation in all districts of the central and peripheral nervous systems (CNS and PNS) where acetylcholine acts as a neurotransmitter.

72.1.2.1  Acetylcholinesterase and Cholinergic Functions AChE is an elongated molecular structure formed by heterologous subunits that is localized mainly in the outer basal lamina of the synapse. The enzyme is highly concentrated at the neuromuscular junction, and it is synthesized in both nerve and muscle. A single gene encodes the catalytic subunits of AChE (Taylor, 1996a). The atomic structure,

determined by x-ray analysis, reveals that the active site catalytic triad (serine, histidine, and glutamate) lies near the bottom of a deep and narrow gorge that reaches halfway into the protein. Substrates and inhibitors are drawn to the active site of the enzyme by an aromatic guidance mechanism that is formed by 14 aromatic residues that line about 40% of the gorge, providing an array of low-affinity binding sites (Sussman et al., 1991). The synaptic function of AChE is to remove the neuro­ transmitter acetylcholine. The hydrolysis of acetylcholine involves acetylation of the serine residue followed by restoration of the active center of AChE. The time required for hydrolysis of acetylcholine at the neuromuscular junction is less than a millisecond (Taylor, 1996a). However, because the cholinergic transmission is involved in a variety of functions that require specific features at different sites, AChE distribution and hydrolysis time may vary, depending on the type of response needed in a given district. Moreover, differences in the cholinergic transmission also depend on several other factors, including synthesis storage and release of acetylcholine, distribution of different receptors, and types of signal transduction (Brown and Taylor, 1996; Taylor, 1996a, b). Cholinergic transmission is involved in the stimulation of skeletal muscle and autonomic ganglia (nicotinic), autonomic effector cells in the white muscles, cardiac conduction system, and secretory glands (muscarinic), and CNS neurons.

72.1.2.2  Chemistry and Biochemistry of Anticholinesterases OPs that display anticholinesterase activity are triesters of phosphoric acid with the general structure shown in Table 72.1 and in Figs. 72.1 and 72.2. R1 and R2 vary

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R1

O

P R2

Figure 72.1  Chemical structure of organophosphorus esters. Oxygen can be replaced with sulfur at all sites. Oxygen atoms bound to R1 and R2 can also be replaced with nitrogen or may be absent on one (phosphonates) or both groups (phosphinates). R groups can be the same or different. Examples of the chemistry of R groups and of leaving groups (X) are also given (also see Table 72.1).

O O

X

O

–CH3 ;–C2H5 ; –IsoC3H7 –F; –CN;

–CH2CH2SC2H5;

– CH –COOC2H5

NO2; – CH = CCI2

– CH – COOC2H5

RO

RNH

O P

P

RO

RNH

X Phosphates

R

S

N

O

R

X

RO

Phosphoroamidates R

O

O

S

X Phosphonates

R

O

P

O

R

X Sulphonates

N-C X

Phosphinates

Figure 72.2  Chemical structure of anticholinesterase agents. Some are pesticides (phosphates, phosphoroamidates, phosphonates and carbamates); others are not (phosphinates and sulfonates). Carbamates are reversible inhibitors of AChE.

O P

R

X Carbamates

O

AChE – OH

+X

O

P

O

O–R

AChE – OH

O-R

+ HO

P

O–R O-R

1 3 O

AChE – OH

X

O

P

O-R O-R

2

O

AChE – O

P

O–R O-R

+X

4 O

AChE – O

P

OO-R

+R

Figure 72.3  General representation of biochemical interactions between OPs and AChE. Reaction 1 leads to the formation of Michaelis complex and reaction 2 leads to phosphorylated AChE. Rates of these reactions indicate the affinity of enzyme for a given OP. Reaction 3 is spontaneous reactivation of AChE, which is usually very slow, although in the case of dimethoxy phosphorylated AChE the speed of reactivation is higher. Reaction 4 leads to a stable, negatively charged phosphorylated AChE (aging of phosphorylated AChE).

(e.g., alcohols, amides) as does the X moiety (e.g., fluorine, phenoxy). The nomenclature of OPs and other anticholinesterases, and their classification may follow various schemes (Chambers, 1992; Edmundson, 1988; Holmstedt, 1963). Figure 72.3 and Table 72.3 illustrate the biochemical

interactions between OPs and AChE, and Table 72.3 offers some examples of them. OPs react covalently with AChE by phosphorylating the serine residue at the catalytic center. This occurs essentially in the same manner that acetylcholine acetylates AChE

Chapter | 72  Clinical Toxicology of Anticholinesterase Agents in Humans

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Table 72.3  Examples of Interactions of Some OPs with Human AChEa Organophosphates

Inhibition

Phosphorylated AChE

Spontaneous reactivation

(AChE I50 106 M) CH3O CH3O

0.95

i-C3H7O i-C3H7O i-C3H7O

O O P O P O O

N

CI CI CI CI

O

E

P O CH = CCI2

CHO C22H33O CHO C22H33O

i-C3H7O

Dichlorvos

O

Clorpyrifos-oxon

0.007

O P

O

E

O P

CI

OP F

DFP

OCH3

OC2H3 OC2H3

O P

O O P O

0.83

P F

O

OCH3

O

E

O P

OC3H7-iO P OC3H7-i

Aging

(t1/2 hours)

(t1/2 hours)

0.85

3.9

OCH3 OCH3

58

41

OC2H3 OC2H3 OCNo 3H7-i reactivation

at 6

4.6

OC3H7-i

a

Data from Lotti and Johnson (1978), Capodicasa et al. (1991), and EPA (1992).

(Aldridge and Reiner, 1972; reactions 1 and 2 in Fig. 72.3). Affinity constants vary, depending on the OP involved. However, in contrast to the acetylated enzyme, which rapidly yields acetic acid and restores the catalytic center, the phosphorylated enzyme is stable and catalytic activity recovers very slowly (reaction 3 in Fig. 72.3). The rate of spontaneous reactivation depends on the chemistry of the attached phosphoryl residue. In the case of dimethoxyphosphorylated AChE, rates are much higher compared with those of phosphoryl residues with longer carbon chains (diethoxy, dipropoxy, etc.), where reactivation might not occur at all. Rates of spontaneous reactivation are even higher for carbamylated AChE, which restores its catalytic activity more rapidly than the phosphorylated enzyme, although still much more slowly than acetylated AChE. For this reason, carbamates are defined as reversible AChE inhibitors. Rates of spontaneous reactivation should be considered in conjunction with those of a further nonenzymatic reaction that occurs on phosphorylated AChE. This reaction, called aging, involves the loss of one alkyl group and leads to stabilization of the phosphorylated enzyme (reaction 4 in Fig. 72.3). The degree of AChE inhibition and its duration in vivo largely depend on these rates: when the rate of spontaneous reactivation is higher than that of the aging reaction, almost complete recovery of activity is expected. On the contrary, if the rate of aging is higher than that of spontaneous reactivation, then irreversible inhibition takes place (Table 72.3). Rates of spontaneous reactivation and aging may have clinical relevance in the case of poisoning by dimethoxy OPs (and by carbamates, where carbamylated AChE does not undergo aging) because diagnostic and therapeutic attitudes may differ from those in cases of poisoning by other OPs (see Section 72.1.3.8).

72.1.3  Clinical Manifestations 72.1.3.1  General Features The clinical pictures of acute OP poisoning reflect the degree of accumulation of neurotransmitter that causes cholinergic over-stimulation in various organs. Early effects are characterized by stimulation or facilitation at various sites, which are followed, at higher concentrations of anticholinesterases, by inhibition or paralysis (Taylor, 1996a). The relationship between OP toxicity and nervous tissue AChE inhibition is influenced by many factors. In general, 50–80% of AChE must be inactivated before symptoms are noted. Brain AChE activity around 10–15% of normal is associated with severe toxicity, and below 10% with coma, seizures, respiratory failure, and death. Lethal exposures in the absence of treatment have been estimated to correspond to approximately 30–50 times the minimal symptomatic exposure for most OPs (Holmstedt, 1959). A pharmacological description of acute poisoning is reported in Table 72.4, where signs and symptoms are ranked as muscarinic, nicotinic, and CNS. Signs and symptoms of acute poisoning usually appear within minutes or a few hours of exposure, depending on the chemical involved, route of exposure, and dose. Unusual cases of suicide attempts by intravenous injections of OPs have been reported, where symptoms and signs appeared quite early even if doses were (probably) small (Güven et al., 1997; Lyon et al., 1987). Cases have been reported of delayed onset of symptoms and signs often associated with prolonged toxicity and relapses. Five patients displayed mild cholinergic toxicity within a few hours of ingestion of dichlofenthion, but severe toxicity did not appear until 40–48 h afterward (Davies et al., 1975). Two patients died and cholinergic

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Table 72.4  Signs and Symptoms of Organophosphate Poisoning Manifestations

Signs and symptoms

Muscarinic Respiratory system

Wheezing, dyspnea, cyanosis, bronchorrhea, bronchospasm, pulmonary edema

Gastrointestinal system

Anorexia, nausea, vomiting, diarrhea, abdominal pain, fecal incontinence

Cardiovascular system

Bradycardia, hypertension

Urinary system

Urinary incontinence

Glands

Hypersalivation, hyperlacrimation, increased sweating

Pupils

Miosis, unreactive to light

Nicotinic Red muscles (including respiratory muscles) Central nervous system

Weakness, fasciculations, twitching, tachycardia, hypertension Headache, drowsiness, dizziness, confusion, blurred vision, slurred speech, ataxia, coma, convulsions, depression and block of respiratory center

symptomatology persisted up to 48 days in the survivors. In one patient, blood dichlofenthion was detected for 75 days after poisoning, whereas in another patient, inhibition of both plasma and red blood cell (RBC) cholinesterases was detected for 66 days. In another case of combined dermal and inhalation exposure to fenitrothion, mild symptoms appeared after 2 days and more severe symptoms appeared over the next 3 days. Relapses of cholinergic signs occurred on days 11 and 17 (Ecobichon et al., 1977). In a case of ingestion of fenitrothion, delayed onset of poisoning was confirmed (Sakamoto et al., 1984). A case of suicide attempt was described where the patient started to complain of symptoms 5 days after ingestion of fenthion and relapse of cholinergic toxicity occurred on day 24 (Merrill and Mihm, 1982). In another suicide attempt with fenthion, the patient had initial symptoms a few minutes after ingestion that became severe 31 h later and lasted for 18 days (Mahieu et al., 1982). In a further case of poisoning by fenthion, blood levels of the chemical measured 11 days after intoxication were 1000 times higher than those measured upon admission to the hospital (Martinez-Chuecos et al., 1992). In a case of suicide attempt with isofenphos,

severe symptoms occurred 24 h after intramuscular injection and lasted for 10 days (Zoppellari et al., 1997). All these cases involved OPs with slow disposal and long persistence in the fat; both these pharmacokinetic characteristics are justified by the high partition coefficient of these compounds (Tomlin, 1997). These chemical and toxicological characteristics of certain OPs should be kept in mind either when patients are first observed or when antidotal treatment is discontinued. The first signs to appear are usually muscarinic, which may or may not be in combination with nicotinic signs. The incidence of signs and symptoms is variable, depending on the dose, the chemical involved, and the time after exposure when detected. Respiratory failure is the hallmark of the clinical picture of severe OP poisoning, whereas mild poisoning and/or early stages of an otherwise severe poisoning may display no clear-cut signs and symptoms. Therefore, diagnosis is made through symptom recognition, followed by grading of poisoning severity, although the latter is only a guide for immediate treatment and has no prognostic value (Bardin et al., 1987; Bardin and van Eeden, 1990; Lotti, 1991; Minton and Murray, 1988). Miosis is observable in more than 80% of patients and in the case of mild poisoning, may represent the only sign (Rengstorff, 1985, 1994). Anorexia, nausea, and vomiting are reported in 40–80% of patients and may also be the only and earliest signs of poisoning (Bardin et al., 1987; Grob and Harvey, 1958; Hayes et al., 1978; Namba et al., 1971; Ohbu et al., 1997; Okumura et al., 1996; Saadeh et al., 1996; Tafuri and Roberts, 1987; Tsao et al., 1990). Diarrhea and abdominal pain are also reported in 20–60% of patients. Hypersalivation is reported in more than 60% and excessive sweating in more than 30% of patients. Hyperlacrymation is less frequent (10–30%). Bronchial hypersecretion and respiratory distress usually follow other muscarinic signs, but not always. Urinary and fecal incontinence can be observed in the most severe cases. Weakness is the only nicotinic symptom that appears at early stages of poisoning. Muscle fasciculations appear later when the clinical picture becomes more severe. CNS signs such as coma and convulsions appear after muscarinic and nicotinic symptoms, although early CNS symptoms may include headache, dizziness, and blurred vision. Based on the recording and evaluation of this constellation of signs and symptoms, and on the circumstantial evidence of exposure, diagnosis is relatively easy. The course of the illness depends on the toxicological characteristics of the OP, on the severity of the clinical picture, and on the promptness and efficacy of treatment.

72.1.3.2  Respiratory Failure As result of combined nicotinic, muscarinic, and CNS cholinergic overstimulation, severe OP poisoning results in respiratory failure. Thus, bronchoconstriction, bronchorrhea,

Chapter | 72  Clinical Toxicology of Anticholinesterase Agents in Humans

pulmonary edema, fasciculations and paralysis of respiratory muscles (the diaphragm in particular), and depression of the brain respiratory center all contribute to respiratory insufficiency. This clinical condition usually develops within 24 h of exposure and within a shorter period from the onset of signs and symptoms. It should be distinguished from a somewhat delayed respiratory failure (one to several days after poisoning) that characterizes the intermediate syndrome, which has a different pathophysiology (see Sections 72.2.2 and 72.2.3.1). The severity of respiratory failure in 52 patients with acute OP poisoning was graded according to the presence/absence of acidosis, which in turn predicted the survival rate, which is much higher in patients with hypoxemia only (Goswamy et al., 1994). In the same study some markers at the time of physical examination were identified as indicative of the need for artificial ventilation. These markers include miosis, hypotension, fasciculations, unconsciousness, and low plasma cholinesterases. As a single sign, fasciculation was the most significant prognostic marker for ventilator requirement and final outcome. This observation indirectly suggests that respiratory muscle failure is more important than any other factor in causing respiratory insufficiency. However, some authors indicated the failure of central respiratory drive as the important factor (Grob, 1956; Tsao et al., 1990; see Section 72.1.3). The use of plasma cholinesterase (ChE) levels to assess poisoning severity and to predict the development of respiratory failure, as also suggested by others (Tsao et al., 1990), should be discouraged because low plasma ChE do not necessarily correlate with the severity of cholinergic overstimulation (see Section 72.1.3.5 and 72.4.3.2).

72.1.3.3  Cardiac Manifestations Given the cholinergic innervation of the heart (Lefkowitz et al., 1996), several types of cardiac alterations are included in the clinical picture of acute poisoning. In a review of 168 cases of acute poisoning that involved a variety of OPs (including dimethoate, methylparathion, trichlorphon, sevin, mevinphos, dichlorvos, and malathion), 134 patients showed electrocardiographic abnormalities, including prolonged QT interval and ST and T abnormalities. Fifty-six patients had arrhythmias 3–15 days after poisoning: ventricular extrasystoles and ventricular tachycardia with “torsade de pointes” characteristics have been observed. Bradyarrhythmias were less frequent, although two cases of AV block occurred (Kiss and Fazekas, 1979). QT prolongation and/or polymorphous ventricular arrhythmias also were observed in 14 patients of another series of 15 cases of OP poisoning that involved phosdrin, parathion, and phosphamidon (Ludomirsky et al., 1982). A case in which sinus bradycardia, AV dissociation, idioventricular rhythm, multiform ventricular extrasystoles, and prolongation of the PR, QRS, and QT intervals

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was observed. Polymorphic ventricular tachycardia, characterized by extreme variability of QRS morphology and changes in the RR interval, was also present (Brill et al., 1984). QT prolongation was recorded in 97 patients in a series of 223 OP poisonings, when electrocardiograms (ECGs) were examined in retrospect (Chuang et al., 1996). Mevinphos, parathion, methamidophos, ethyl 4-nitrophenyl phenylphosphonothioate (EPN), and other unspecified OPs were involved. The authors concluded that patients who presented with such ECG changes have higher mortality and a higher incidence of respiratory failure compared with those without QT prolongation. These patients were also among those who had the highest plasma cholinesterase inhibition. However, whereas patient ranking was based on plasma cholinesterase levels, and QT changes and respiratory failure also were detected in patients graded as having very mild to moderate poisoning, it is not possible to appreciate the prognostic value, if any, of these ECG changes. In another review of 46 cases of OP and carbamate poisoning, 31 patients developed QT prolongation with or without ST and T changes, among which, 4 had ventricular tachycardia and 2 had ventricular fibrillation. Sinus tachycardia and bradycardia were equally represented (Saadeh et al., 1997). In 50 fatal cases of OP intoxication, focal myocardial damage was observed at autopsy, including pericapillar hemorrhage, micronecrosis, and patchy fibrosis (Kiss and Fazekas, 1982). Post mortem ultrastructural examination of the heart was also performed on 10 patients who died of acute poisoning by azinphos-ethyl (9) or dimethoate (Pimentel and Carrington da Costa, 1992). Patients died 3–17 days after poisoning, all having shown ECG changes during the illness. Lysis of myofibrils, swollen and fragmented mitocondria, disorganization of nuclear chromatin, and Z band abnormalities were observed. Whether these changes are primarily related to cholinergic overstimulation of the heart and subsequent arrhythmias or are secondary to the general condition of the patients cannot be ascertained. In conclusion, several types of arrhythmias may develop during acute OP poisoning. Therefore, patients must be carefully monitored and promptly treated.

72.1.3.4  Central Nervous System Manifestations Various manifestations of CNS involvement that grossly reflect the severity of poisoning have been described. Thus symptoms and signs range from headache, anxiety, confusion, sleep disturbances, and blurred vision to tremor and convulsions to coma, hypothermia, and central respiratory depression. Some authors believe the failure of the central respiratory drive is an important factor in causing respiratory

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failure (Grob, 1956; Tsao et al., 1990), although clinical discrimination of the selective effects of anticholinesterases on the CNS is difficult, particularly in assessing their relative importance in causing death (Lotti, 1992a), because toxicity to the CNS, in general (Norton, 1986), and of anticholinesterases, in particular (Glow and Rose, 1965), always holds a strong peripheral component (see also Section 72.1.3.2). Early CNS signs and symptoms may last for several days, in which only moderate additional cholinergic symptomatology may be detectable (Grob and Harvey, 1958). Electroencephalogram (EEG) abnormalities can be detected at the onset of symptoms and are characterized by irregularities in rhythm, variation and increase in potential, and intermittent bursts of abnormally slow waves of elevated voltage similar to those seen in epilepsy; these symptoms usually persist for about a week or longer (Grob et al., 1947). Coma is usually due to direct CNS depression by OPs, although hypoxia that derives from respiratory failure may contribute. In this condition, EEGs show profound depression of cortical activity (Lotti and Becker, 1982a). OPs vary in their potency to induce seizures (Hoskins et al., 1986) and perhaps this manifestation is not entirely due to AChE inhibition (Van Meter et al., 1978), considering that it is blocked by benzodiazepines, which is known to act via -aminobutyric acidergic (GABAergic) mechanisms (Lipp, 1973; see also Section 72.1.3.8). Occasionally, during the acute phase of OP intoxication, patients displayed opsoclonus (De Bleecker, 1992; Hata et al., 1986; Pullicino and Aquilina, 1989). Opsoclonus is an abnormal, rapid, involuntary, repetitive, chaotic, and conjugated ocular movement. Different OPs were involved in reported cases; the onset was from 3 to 48 h and there no was correlation with the typical symptomatology, although all patients were severely poisoned. Opsoclonus is most likely another sign of cholinergic overstimulation, which lasts for several hours to a few days and recovers spontaneously. One case of poisoning with chlorpyrifos and two others with unknown OPs presented with choreo-athetosis (Joubert et al., 1984; Joubert and Joubert, 1988). Extrapyramidal signs lasted a few days during recovery from acute intoxication. Similar signs also have been reported several days after the onset of cholinergic toxicity (see Section 72.1.3.6).

72.1.3.5  Laboratory Findings Red Blood Cell and Plasma Cholinesterases In addition to synapses, AChE is also present in the outer membrane of red blood cells and, to a lesser extent, in plasma. Its physiological functions in blood are unknown. Plasma butyrylcholinesterase (BuChE), also known as pseudocholinesterase, has a different substrate specificity because it hydrolyzes butyrylcholine. Its physiological functions also are not known in plasma or elsewhere.

BuChE inhibition by OPs is, therefore, not necessarily indicative of exposures high enough to cause poisoning. Moreover, certain diseases and genetic conditions are characterized by low levels of plasma BuChE (see Section 72.4.3.2). Several methods are available to measure these blood enzymes (among the many are Doctor et al., 1987; Ellman et al., 1961; Garcia-Lopez and Monteoliva, 1988; Lewis et al., 1981; London et al., 1995; St. Omer and Rottinghaus, 1992; Wilson et al., 1996). However, because hospital laboratories rarely measure RBC AChE activity, in most circumstances one should rely on measurements of plasma BuChE, for which kits are easily available. RBC AChE inhibition confirms the diagnosis of acute OP poisoning. Whole blood AChE also may be measured, considering that only about 10% of the activity is due to the plasma enzyme (Worek et al., 1999b). Usually there is a good correlation between the severity of signs and symptoms of poisoning and the degree of inhibition of RBC AChE (Table 72.5). Nevertheless, because acute poisoning usually requires prompt treatment, treatment should not be delayed while laboratory confirmation is sought. Therefore, measurement of RBC AChE has limited value in an emergency because diagnosis is exclusively clinical and severe poisoning is inevitably associated with high RBC AChE inhibition. It is more difficult to interpret the relatively low levels of RBC and plasma cholinesterases such as those observed in cases of poisoning that present with equivocal symptoms. This may be the case in mild poisoning or in the initial phase of poisoning caused by OPs that are slowly disposed. Reasons for these difficulties are manyfold and include the following: Large inter- and intraindividual variability of both RBC AChE and plasma BuChE, making the distinction between physiologically low and inhibited activities impossible. Methods to address these difficulties have recently been developed but they have not been applied yet in clinical settings (Polhuijs et al., 1997; see Section 72.4.3.2). ● Different sensitivity of AChE and BuChE to the same inhibitor. In many cases, the identity of the chemical involved and knowledge of its biochemical characteristics are unknown, thus hampering the interpretation of a significant inhibition of plasma BuChE associated with little or no inhibition of RBC AChE. This may be the case in both a mild poisoning or the early phase of a more severe poisoning when poisonings are caused by OPs that preferentially inhibit BuChE (see Section 72.4.3.2). ● The ratio between the inhibition of RBC AChE and that in the synapses may vary according to the compound. Inhibition of the RBC enzyme may be detected without clinical signs of toxicity in cases of exposures to OPs that do not easily cross the blood-brain barrier (see Section 72.4.3.2). ●

Chapter | 72  Clinical Toxicology of Anticholinesterase Agents in Humans

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Table 72.5  Correlation between Severity of Poisoning, Inhibition of RBC AChE, and Symptoms and Signs of OP Poisoninga Severity of poisoning (approximate activity of RBC AChE)

Symptoms and signs Muscarinic

Nicotinic

Central nervous system

Mild (RBC AChE  40%)

Nausea, vomiting, diarrhea, salivation/lachrymation, miosis bronchoconstriction, increased bronchial secretions, bradycardia

Usually none

Headache, dizziness

Moderate (20%  RBC AChE  40%)

Same as above plus pupils unreactive to light, urinary/ fecal incontinence

Muscle fasciculation (fine muscles)

Same as above plus dysarthria, ataxia

Same as above Same as above plus muscle fasciculation (diaphragm and respiratory muscles)

Same as above plus coma, convulsions

Severe (RBC AChE  20%)

a

Modified from Lotti (1991).

If reactivators (oximes) are administered, the pharmacological effect depends on the ratio between inhibited and aged enzyme. This ratio may be different in blood and in the nervous tissue, and the pharmacological reactivation of inhibited blood enzymes will be more effective than that of enzymes in the nervous system. Under these circumstances relatively small inhibition in blood enzymes would not correlate with symptomatology (see Section 72.4.3.2). ● The method of cholinesterase measurement involves dilutions, which in the case of inhibitors such as carbamates and dimethyl phosphates, may favor the spontanous reactivation of inhibited enzymes (see Section 72.1.2.2). These assay related problems would understimate the actual in vivo inhibition. ●

In conclusion, measurements of blood cholinesterases may confirm the diagnosis, but are not essential: clinical observation remains the cornerstone for diagnosis. For the same reasons, repeated measurements of blood cholinesterase during poisoning have no prognostic value (Nouira et al., 1994) and they cannot be used to assess the efficacy of treatment. When AChE is irreversibly inhibited by OPs, the reappearance of RBC activity depends on new erythrocytes entering the blood stream. Whereas the average lifespan of RBCs is 120 days, in most cases, observed reappearance of RBC AChE occurs at a rate of about 1% per day. The corresponding rate for plasma BuChE, which derives from liver synthesis, is about 5% per day (see Section 72.4.3.2). Measurements of OPs and Metabolites in the Blood and Urine  Several analytical methods based on chromatographic techniques are available for quantitative and qualitative measurements of OPs and their metabolites in body fluids (Tomlin, 1997). These measurements have

confirmatory uses in clinical toxicology although they are rarely performed because they are not easily available. Nevertheless, serial measurements of the parent compound in blood (Fig. 72.4) identifies the compound(s) and provides information about its pharmacokinetics in humans. Metabolites of OPs can be measured in the urine. They are products of hydrolytic reactions, and include the alkylphosphate and the alcoholic moieties. Because the kinetics of urinary excretion of metabolites differ, quantitative extrapolations are almost impossible unless complete urine collection is performed over several days and both metabolites are measured. Moreover, if only one of these metabolites is measured, the results may be less specific because different OPs can share identical acidic or alcoholic moieties. Thus, two parent OP compounds that display quite different acute toxicity may produce the same amount of the same metabolite, thereby further hampering quantitative extrapolations (see Section 72.4.3.1). In conclusion, measurements of OPs and their metabolites have a limited value for diagnosis of acute OP poisoning. They may be useful for confirmatory purposes and research. Routine Hematological and Biochemical Tests  Abnor­ malities of almost all common laboratory tests have been reported during the course of acute OP poisoning. None of them is specific and they may reflect either somewhat typical short-term complications (pancreatitis and myopathy; see Section 72.1.3.6) or the general clinical conditions of the patient (degree of respiratory failure, changes in organ perfusion, concurrent infections, iatrogenic consequences, etc.). Isolated reports indicate that activation of blood coagulation was observed in cases of parathion and dimethoate poisoning, and in one case the patient was treated with eparine (Jastrzebski et al., 1994; Kaulla and Holmes, 1961). However, when studies were performed

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Figure 72.4  Time course of blood concentration of several OPs in acutely poisoned patients. Data for () methamidophos, () fenitrothion, () methylparathion, and () parathion from (Lotti, 2000); () chlorpyrifos from (Lotti et al., 1986); () methidation, () dimethoate, and () mecarbam from (Tsatsakis et al., 1996); malathion from (Lyon et al., 1987); (*) trichlorfon from (Nordgren et al., 1980).

µg OP/mI plasma

100

10

1

0.1

0.01

0

2

4

6

8

10

DAYS

on a series of 31 patients moderately poisoned with either parathion or sarin, both hyper- and hypocoagulability were observed (Kaulla and Holmes, 1961). The clinical and toxicological relevance of these isolated observations remains unclear. In conclusion, changes in common laboratory tests are not specific for the diagnosis of OP poisoning. Electrophysiology  The electrophysiological consequences of AChE inhibition at the neuromuscular junction have been reviewed (Singh et al., 1998b), and are characterized by single electrical stimulus-induced repetitive compound muscle action potentials (CMAPs) and, in response to repetitive nerve stimulation, by a decrement of the CMAP (Besser et al., 1989a, b; Maselli et al., 1986). These events reflect the excess of acetylcholine at synaptic levels and subsequent alteration of nicotinic receptor responses. The repetitive muscle response to a single nerve stimulus is thought to result from reexcitation of the muscle by the prolonged endplate potentials brought about by the excess of acetylcholine. At relatively low AChE inhibition, repetitive stimulation causes a decrement followed by complete recuperation of CMAP amplitudes. However, at higher AChE inhibition, repetitive stimulation results in an unimodal pattern of progressive decrements of CMAP amplitudes. Experimental studies demonstrated that in the former case, impaired neuromuscular transmission is caused by transient depolarization of the endplate region, whereas in the latter, direct blockade is due to desensitization of postsynaptic nicotinic receptors (Maselli and Leung, 1993; Maselli and Soliven, 1991). Decremental response to repetitive nerve stimulation dramatically worsened after injection of edrophonium (Maselli et al., 1986), indicating the effects of a further excess of acetylcholine, whereas the administration of pancuronium improved neuromuscular transmission, indicating a blockade of nicotinic receptors (Besser et al., 1990, 1991). Distinction between these findings and those

detected during the intermediate syndrome may be difficult (see Section 72.2.3.2). Other Tests  On a small series of patients poisoned with OPs, cerebral perfusion was investigated by brain single photon emission computerized tomography (Yilmazalar and Özyurt, 1997). The authors concluded that in severe cases of poisoning, patients showed perfusion defects, especially in the parietal lobe. This study is difficult to assess because some patients who had normal plasma cholinesterase and some mild symptoms probably were not poisoned. Moreover, the patients who had major perfusion deficits were the oldest and almost no improvement was detected 3 months after poisoning. Pathology  Gross pathology and histological examination performed at autopsy after fatal poisoning were either unremarkable or nonspecific (Maresch, 1957). In particular, given the mechanism by which OPs cause death, neuropathology is not observed unless severe hypoxia or convulsions occurred (McLeod, 1985). A common nonspecific feature of CNS histopathology is vascular damage associated with increased permeability of the vessel walls, suggesting major changes in the blood-brain barrier that might occur during OP poisoning (see Section 72.1.3.8).

72.1.3.6  Short-Term Complications Neurological  Fatal encephalopathy was reported in two cases of acute OP poisoning (de Reuck et al., 1979). These patients, after an initial recovery from cholinergic toxicity, developed a severe encephalopathy within 3–4 days and died 9–20 days after admission. Necroscopy showed severe hemorrhagic necrosis of the ventricles, resembling lesions observed in Wernicke’s encephalopathy. There is no explanation for such findings, although it was reported that one patient had a history of alcoholism and the other of recurrent depression.

Chapter | 72  Clinical Toxicology of Anticholinesterase Agents in Humans

Neuroleptic malignant-like syndromes following organophosphate poisoning have been reported in the Japanese literature (Ochi et al., 1995). In one case, a 60-year-old schizophrenic patient who had undergone a frontal lobotomy at the age of 20, but was free of medications since then, attempted suicide by ingesting a large amount of methidathion. About a week after full recovery from symptoms and signs of OP poisoning, the patient developed a high fever, extrapyramidal rigidity, and coma. Serum CPK and LDH were increased, and high urinary myoglobin was detected, whereas plasma cholinesterase was back to normal. The patient was treated successfully with dandrolene. A correlation with OP poisoning is difficult to ascertain because this clinical picture is also consistent with lethal catatonia, a rare and potentially lethal syndrome that occurs in schizophrenic patients (Castillo et al., 1989). Extrapyramidal manifestations that complicate OP poisoning were described in six patients poisoned with fenthion (Senanayake and Sanmuganathan, 1995). Patients had moderate to severe cholinergic signs and all developed intermediate syndrome subsequently. Dystonia was observed in all patients, and the most common signs were tremor, choreo-athetosis, and cog-wheel rigidity. Onset of these extrapyramidal signs was variable (4–40 days) and spontaneous recovery was observed within 4 weeks. Although the development of extrapyramidal signs is rarely described in OP poisoned patients (and in these cases, it may be compound specific), a possible relationship with inadequate oxime therapy has been inferred. Transient bilateral vocal cord paralysis has been reported after OP poisoning, although the reports are inconsistent and describe different clinical characteristics. A 3-year-old boy severely poisoned with chlorpyrifos was intubated and artificially ventilated (Aiuto et al., 1993). On day 3, immediately after extubation, the boy developed stridor and was reintubated. He was reextubated on day 6 and specific treatment for OP poisoning was halted. Occasional stridor was treated conservatively until day 11, when he needed intubation again. Tracheostomy was performed on day 19 and maintained for more than 50 days. Generalized areflexia was observed on day 18 that rapidly recovered within a week. Follow-up was not reported. Although the authors suggested that organophosphateinduced delayed polyneuropathy (OPIDP) developed, the distribution of weakness and the time course of events is not characteristic (see Section 72.3.3.1). Three cases of delayed recurrent laryngeal nerve paralysis (25–35 days from the onset poisoning) were observed in subjects severely poisoned with three different OPs (chlorpyrifos, methamidophos, and parathion) who required intubation and artificial ventilation (de Silva et al., 1994). At the onset of stridor, no other nerve was involved and the patients had been extubated 14–26 days earlier. One patient required tracheostomy, whereas the others were treated conservatively. Recovery occurred within

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4–15 weeks. Although OPIDP was suggested as the cause of laryngeal nerve paralysis, the isolated involvement of this nerve speaks against it. Moreover, parathion is not known to cause OPIDP (see Section 72.3.1). Another case of isolated bilateral vocal cord paralysis was associated with OP exposure (Thompson and Stocks, 1997). However, exposure was anecdotal and there was no evidence of OP poisoning. In further case of poisoning with unspecified OPs, bilateral vocal fold palsy was reported (Indudharan et al., 1998). Palsy was observed at extubation, about 10 days after poisoning, and required tracheostomy that was maintained for 2 months. The authors suggested intermediate syndrome, but it was unknown when, after poisoning, this palsy occurred (because the patient was intubated) and if other signs of proximal weakness were detected (see Section 72.2.3.1). Cortical visual loss was observed in two patients poisoned with OPs (Wang et al., 1999). One patient had cortical visual loss as shown with positron emission tomography (PET) scan 1 year after poisoning with EPN. During this period, blurred vision was the main symptom and was not clear how long she was hypoxic. Another patient had similar effects that were detected 2 months after acute poisoning with mevinphos. She was reported to have had apnea, but the duration was unspecified. Therefore, it is likely that these two cases represent consequences of secondary hypoxia, although selectivity for this cortical area is unusual. A case of moderate poisoning with a mixture of di­methoate, diazinon, and methoxychlor was reported where the patient was markedly hypothermic 1 h after ingestion in addition to having miosis and diffuse fasciculations. She had a rectal temperature of 33°C, which normalized within 1 h after passive rewarming (Hantson et al., 1996). In conclusion, there is no strong evidence for any neurological short-term complication directly linked to OP toxicity, except, perhaps, hypothermia, which has been consistently reported in animal studies as a compound-specific effect of certain OPs (Coudray-Lucas et al., 1983). Pancreatitis  Although it is uncommon, pancreatitis has been consistently associated with acute OP poisoning. Nevertheless, because the diagnosis of pancreatitis is usually established by detection of an increased level of serum amylase, quite often the roles of salivary gland overstimulation and/or of acidosis, both present in OP poisoning, have been overlooked as a cause of increased serum amylase. Additionally, in most reports there is no information on possible alternative causes of pancreatitis, including anoxia, hypoperfusion, infections, and concurrent treatment. The latter is important for a differential diagnosis of toxic pancreatitis, because about 5% of causes of acute pancreatitis are related to commonly used drugs (Greenberger et al., 1998). Transient hyperglycemia and glycosuria, which are often found in severe OP poisoning

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(Namba et al., 1971; Zadik et al., 1983), are also common in acute pancreatitis. Hypotheses have been formulated that link the pathophysiology of the triad of pancreatitis, hyperamylasemia, and hyperglycemia in acute and severe OP poisoning (Haubenstock et al., 1983). The first case was reported on a patient who was anoxic for an unspecified period, was misdiagnosed as acute kerosene inhalation, and was later found to be intoxicated with OPs. High serum amylase levels were almost immediately detected and after 9 days, a pancreatic pseudocyst was drained (Dressel et al., 1979). A case of OP poisoning by coumaphos was reported where the patient with severe respiratory failure had elevated serum amylase on admission that returned to normal 20 h later (Moore and James, 1981). Another case of pancreatitis was less convincingly attributed to cutaneous exposure to an organophosphate insecticide. Symptomatology was mild though consistent with OP poisoning but there was no evidence of poisoning and medical history suggested a very mild exposure, if any. Symptoms of pancreatitis persisted for 6 months (Marsh et al., 1988). Acute painless pancreatitis developed in a patient soon after severe intoxication with mevinphos (Hsiao et al., 1996). RBC AChE was inhibited, serum amylase and lipase were elevated, and a computed axial tomography (CT) scan indicated pancreatitis. Persistent hyperglycemia developed. Two cases of acute severe pancreatitis, complicated by pancreatic necrosis and retroperitoneal sepsis were described in patients with severe poisoning by unspecified OPs (Panieri et al., 1997). In both cases, diagnosis of pancreatitis was made several weeks after admission (2 and 5, respectively) and confirmed at surgery. On a series of 75 patients admitted to the hospital for malathion poisoning, serum amylase was serially measured (Dagli and Shaikh, 1983): 47 patients had a mildly raised amylase that reversed within 2 days. Mild symptoms compatible with OP poisoning were reported in most patients, but it is not clear if they were the same patients with elevated amylase. Apparently none had severe poisoning and no toxicological evidence of OP poisoning was provided. In another series of nine patients poisoned with parathion, painless acute hemorrhagic pancreatitis was manifested by ileus in two cases (Lankisch et al., 1990). Patients were severely ill, requiring artificial ventilation, and had depressed plasma cholinesterase levels. One subject developed persistent ileus 1 week after admission and hemorrhagic pancreatitis was observed at surgery. The other patient presented with severe shock and paralytic ileus. Blood lipase and amylase were elevated and the CT scan indicated pancreatitis. A series of 17 children poisoned with OPs was compared with a matched control group with similar abdominal symptoms (Weizman and Sofer, 1992). Compounds were identified in nine cases and included parathion, malathion, and diazinon. Five poisoned children were diagnosed with acute pancreatitis based on elevated serum amylase trypsin

Hayes’ Handbook of Pesticide Toxicology

and glucose. Ultrasonography was not performed. None of the controls developed pancreatitis. It is not clear when pancreatitis developed after exposure, if adominal pain in these patients was, due to either pancreatitis or cholinergic overstimulation, how severe the poisoning was and what disease was diagnosed in controls. In a retrospective study on 32 cases of OP poisoning (unspecified chemicals), 16 patients developed respiratory failure (Matsumiya et al., 1996). The average levels of plasma amylase of these patients on admission was higher than that of patients who did not developed respiratory failure. The authors concluded that high amylase was predictive of subsequent respiratory failure. However, several patients with respiratory failure had normal plasma amylase and several patients without respiratory failure had elevated plasma amylase. In a retrospective study on 159 patients admitted to the hospital for OP poisoning, Lee et al., (1998) found hyperamylasemia in 44 patients (out of 121 for which data were available) and hyperlypasemia in 9 (out of 28). Nine patients were diagnosed as having pancreatitis and two of them died. The incidence of hyperamylasemia was related to the clinical severity of poisoning. It is difficult to conclude from the described clinical data whether OPs are directly toxic to the pancreas. Different OPs have been involved in these cases, the onset of pancreatitis seems to be quite variable after the beginning of cholinergic symptomatology, and important details are missing in many reports. Moreover, there are indications that in most circumstances, shock and subsequent hypoperfusion preceded the development of pancreatitis. In conclusion, although pancreatitis is consistently reported, there is no strong evidence that it is a characteristic feature of acute OP toxicity. Myopathy  On several occasions, junctional myopathy, that is, muscle damage that originates in the endplate region, has been observed in poisoned patients. The first description of muscle fiber necrosis was obtained from a case of severe parathion poisoning (de Reuck and Willems, 1975). The patient was admitted with severe respiratory failure and was artificially ventilated. Myoclonic jerks and fasciculations were present and blood cholinesterases were profoundly depressed. Episodes of irregular tachycardia developed on day 4 and the patient died of cardiac arrest on day 9. Post mortem examination revealed patchy and focal areas of necrosis of muscle fibers in the diaphragm. Nerve endings in the segmental necrotic zones of the muscle fibers were degenerated, whereas normal motor endplates were observed in the nonnecrotic muscle fibers. Similar lesions involving muscles other than the diaphragm were reported after acute poisoning with trichlornate (de Reuck et al., 1979). The patient had moderate signs of acute OP poisoning including muscle fasciculations. After an initial improvement, the patient became comatose and required artificial ventilation on day 6, and eventually died 20 days after admission because of Gram-negative

Chapter | 72  Clinical Toxicology of Anticholinesterase Agents in Humans

sepsis. Post mortem examination revealed waxy degeneration and lysis of individual fibers in the quadriceps femori muscle and, to a lesser degree, in the deltoid, diaphragm, and intercostal muscles. Nerve fibers were normal. In another study, samples of intercostal muscles were obtained at autopsy from a subject who died after an unquantified exposure to malathion and diazinon. Toxicological evidence of exposure and information on the severity of cholinergic overstimulation were scanty. Moreover, the patient died of brain hemorrhage. Histopathology of muscles showed basophilic inclusions and scattered necrotic fibers. It is difficult to attribute these changes to OP exposures and even more difficult to attributed them to prolonged AChE inhibition. In fact, muscular AChE was inhibited to about 50% and plasma cholinesterases were inhibited perhaps a little more (Wecker et al., 1985). Another case report associated myopathy with OP exposure, although it was much less convincing (Ahlgren et al., 1979). The patient, who was working as an exterminator, had never had acute poisoning. Sometimes he exhibited symptoms compatible with cholinergic overstimulation, but they rapidly subsided. He presented with a 2-year history of muscle weakness, involving mainly the trunk, the shoulders and the pelvic girdle. Several muscle biopsies performed during the course of the disease revealed necrotic fibers, inflammatory reactions, and fibrotic changes in the deltoid muscle and quadriceps femoris muscles. Five years after the onset of symptoms, this patient died of progressive respiratory failure. At autopsy generalized atrophy and fibrosis of striated muscles was observed; and in particular, the diaphragm was almost entirely fibrotic. It is difficult to judge whether chronic exposure to OP caused the progressive myopathy. Clinical expression of any toxic disease is usually complete after cessation of exposure, and this patient reduced and eventually halted the exposure because of the clinical condition. No other reports indicate the progression of acute myopathy beyond the initial period of cholinergic overstimulation. There is no quantitative evidence of exposure, which was presumably low and comprised several pesticides, although the patient reported that he most frequently sprayed diazinon. In conclusion, there is limited evidence that acute OP poisoning causes myopathy in humans and it may be the consequence, in the cases described, of either fasciculations or sepsis. Muscle necrosis, the human equivalent of a well known effect observed in experimental animals (Dettbarn, 1992), was thought to be a possible cause of intermediate syndrome (Karalliedde and Henry, 1993; Senanayake and Karalliedde, 1987; see Section 72.2.2). However, although clinical signs suggestive of red muscle damage (such as elevated CPK) have been observed in some patients who displayed signs of intermediate syndrome (He et al., 1998), muscle biopsies suggest that the lesions were too sparse to justify clinically detectable muscle weakness (De Bleecker, 1993; De Bleecker et al., 1993).

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Others  Severe intoxications carry obvious risks of several secondary manifestations. Examples include a case of diazinon poisoning characterized by loss of fluids complicated with acute renal failure (Abend et al., 1994), a case of acute dimethoate poisoning complicated with Gram-negative pneumonia, acute respiratory distress syndrome, and acute renal failure (Betrosian et al., 1995), and two fatal cases of poisoning with diazinon presenting with severe hyperglycemia, metabolic acidosis, and hypoalkalemia (Hui, 1983). Preexisting diseases may also be worsened by OP poisoning as described in cases of cardiovascular disorders, which have precipitated cerebral infarction and gangrene (Buckley et al., 1994). All these and other complications should be expected in any severe case of poisoning treated in an intensive care unit. The low levels of plasma cholinesterase caused by OP exposure change the pharmacokinetics of drugs that are substrates for these enzymes. Thus, neuromuscular blocking agents, which are widely used in anesthesia, will be slowly metabolized, thereby prolonging their pharmacological action (Øster-gaard et al., 1992). The first case was described in a patient severely poisoned with parathion who had undetectable plasma cholinesterase activity. In an effort to control the patient’s convulsions, small doses of succinylcholine were administered (Quinby et al., 1963). The patient became suddenly apneic and completely flaccid, and the marked neuromuscular block lasted for about 2 h. A patient suspected of having a partial obstruction of the small intestine underwent emergency laparotomy and the anesthesia was induced with thiopentone and suxamethonium (Gesztes, 1966). At inspection, no obstruction was observed and the patient had a prolonged apnea when treatment was discontinued. It was later recognized that the patient had been treated with eye drops that contained ecothiopate iodide—an anticholinesterase drug—that was systemically absorbed. This clinical case is fully justified by the toxicity of ecothiopate. Thus, the pseudo-obstruction associated with the diarrhea was likely a sign of cholinergic overstimulation by ecothiopate and the prolonged apnea was the consequence of plasma cholinesterase inhibition, eventually leading to reduced metabolism of the suxamethonium. A similar case was reported in a girl severely poisoned with chlorpyrifos following administration of succinylcholine for airway management (Selden and Curry, 1987). A patient exposed to malathion underwent surgery for acute appendicitis. Because he had no signs of cholinergic overstimulation, suxamethonium was given to facilitate tracheal intubation (Guillermo et al., 1988). He suffered prolonged apnea due to low plasma cholinesterase activity, which most likely was due to relatively low exposure to malathion because he had a normal plasma cholinesterase phenotype. Another case of a boy who suffered relatively mild exposure to chlorpyrifos and propetamphos, and exhibited no symptoms of cholinergic overstimulation, showed

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prolonged apnea following suxamethonium treatment (Weeks and Ford, 1989). These cases represent one of several possible conditions in which low plasma cholinesterase activity prolongs the pharmacological effects of neuromuscular blocking agents (Davies and Landy, 1998; Hart et al., 1995; Kopman et al., 1978; see Sections 72.4.3.2 and 72.4.3.3). OP intoxication seems to have no effect on pregnancy if patients are properly treated. Pregnant women, between 9 and 36 weeks of gestation, intoxicated with OPs (sarin, methamidophos and fenthion) who received appropriate management completely recovered from the poisoning, allowing the pregnancies to continue to term unaffected and the delivery of healthy babies (Karalliedde et al., 1988; Ohbu et al., 1997).

72.1.3.7  Differential Diagnosis Needless to say, it is vital to distinguish OP poisoning from other diseases, particularly when the differential diagnosis may be hampered by unusual scenarios. Thus, when observing a severely ill patient, an exposure to OPs may not be suspected at first glance (Björnsdóttir and Smith, 1999). Conversely, patients known to have been exposed to OPs may present with illnesses that are unrelated to exposure. Although diagnosis of acute OP poisoning is straightforward, difficulties have been reported. In 20 children transferred from other hospitals, correct diagnosis of OP poisoning was made in only 4, whereas the others had diagnoses of pneumonia, various infectious diseases, encephalopathy and head trauma (Zwiener and Ginsburg, 1988). Conversely, in another series of 78 patients admitted to the hospital with diagnoses of pesticide poisoning (among which 34 were thought to be caused by OPs), only 36 (among which 18 were due to OPs) were confirmed as clinical poisoning (Lamminpää and Riihimäki, 1992). Reasons for misdiagnosis included other illnesses, no evidence of exposure, or proven limited absorption. Although the signs of cholinergic overstimulation are characteristic, they must be actively sought when OP poisoning is suspected, particularly in mild cases, because they are rarely isolated. Detection of concurrent signs would help, for instance, to ascertain the toxic nature of respiratory failure, which should be distinguished from other respiratory and circulatory causes and, similarly, for differential diagnosis of the most serious CNS manifestation such as seizures and coma (Greenaway and Orr, 1996; Hollis, 1999).

72.1.3.8  Treatment Prompt treatment of OP poisoning is lifesaving. Several procedures are available, the sequence of which depends solely on the severity of poisoning. Special attention should be exercised by medical personnel caring for these patients, because passive contamination may occur (Ohbu et al., 1997).

Hayes’ Handbook of Pesticide Toxicology

Minimizing Further Exposure  Procedures aimed at decontamination and/or at minimizing absorption depend on the route of exposure. Thus, in the case of dermal exposure, contaminated clothing should be removed and the skin should be washed with alkaline soap. When the skin appears to be clear, the patient should be bathed or swabbed, because most OPs are more soluble in alcohol than in water (Durham and Hayes, 1962). In the case of eye contamination, extensive irrigation with water or saline should be performed for several minutes. In case of ingestion, various procedures have been recommended to reduce absorption from the gastrointestinal tract, although there is limited evidence of their efficacy (Johnson and Vale, 1992; Lotti, 1991). In conscious patients, vomiting is usually induced with syrup of ipecacuanha (10–30 ml followed by 2–300 ml of water). This treatment is contraindicated if the patient is semiconscious, has difficulty swallowing, or if the pesticide is dissolved in hydrocarbon solvents or is corrosive, given the high incidence of pneumonitis/atelectasias due to insecticides that contain petroleum distillate (Zwiener and Ginsburg, 1988). In these circumstances, the increased probability of aspiration pneumonia largely exceeds that of potential benefits. Gastric lavage with instillation of activated charcoal should be performed after the patient’s airway has been protected with an endotracheal tube. It has been suggested that this procedure be repeated every few hours, as long as the chemical is detectable in the lavage fluid, because it may persist for several days (Futagami et al., 1995; Willems, 1981). Although the value of this continuous lavage is not proven, empirical evidence and the slow absorption characteristic of some OPs (see Section 72.1.3.1) suggest this option be considered. Moreover, complications due to administration of activated charcoal, in addition to aspiration, may also include intestinal ulceration and massive bleeding due to constipation (Mizutani et al., 1991). In fact, when these patients are treated with atropine concurrently, they display decreased bowel peristalsis. Thus, although the use of vigorous cathartic treatment has been suggested as an adjunct to charcoal to hasten the elimination of the charcoal-poison complex, studies in humans have failed to demonstrate any substantial benefit from this combination (Neuvonen and Olkkola, 1988). Other procedures aimed at removing the absorbed OP, for example, hemoperfusion or hemodialysis, have been suggested (Luzhnikov et al., 1977; Verpooten and De Broe, 1984; Yokoyama et al., 1995), but it is doubtful that these methods achieve higher clearance of OPs from the blood than that which occurs spontaneously (Nagler et al., 1981). Moreover, because of the high reactivity and large volume of distribution of OPs, even a high blood clearance will not significantly reduce the total amount of the compound. In a retrospective study on the effects of hemoperfusion, it was shown that this procedure did not remove significant amounts of various insecticides

Chapter | 72  Clinical Toxicology of Anticholinesterase Agents in Humans

(fenitrothion, methylparathion, fenthion, chlorpyrifos, trichlorphon, diazinon, omethoate and fenamiphos) from the body of 10 patients (Martinez-Chuecos et al., 1992). High plasma concentration of OPs reappeared after the last hemoperfusion, symptoms and signs did not ameliorate, and prolonged clinical course and complications were not avoided. Similar results were reported in a fatal case of fenitrothion poisoning treated with combined hemoperfusion and hemodialysis (Yoshida et al., 1987). Atropine  Atropine represents the cornerstone of the treatment for poisoning by anticholinesterase. Other available drugs, such as reactivators of inhibited AChE, which are very effective, should be considered, in principle, as valid adjuncts to atropine administration. Atropine is a muscarinic receptor antagonist that prevents the effects of acetylcholine by blocking its binding to muscarinic cholinergic receptors (Brown and Taylor, 1996). It is a racemic mixture of active 1-hysocyamine and inactive d-hysocyamine that should be stored at 15–30°C and protected from light. Freezing should be avoided. The shelf life is 24 months from the date of manufacture if it is kept under the recommended conditions (Heath and Meredith, 1992). Pharmacokinetic data about atropine are limited. The kinetics of distribution of atropine seem to be dose dependent: about 20% of the drug is bound in plasma and two phases, with apparent half-lives of 1 and 140 min, respectively, have been identified after intravenous injection (Hinderling et al., 1985). However, for practical purposes, the reported plasma half-lives after both intravenous and intramuscular injections, varying between 1.3 and 4.3 h, should be considered. Differences are due to assay methods and to a considerable intra- and interindividual variability (Adams et al., 1982; Kanto and Klotz, 1988; Kentala et al., 1990). In children and in the elderly, the plasma half-life may be longer. The reported apparent volume of distribution is quite large (2–3.5 liter kg1), implying significant intra- and extracellular binding and partitioning of the drug. In children, higher volumes of distribution than in adults have been reported. About 50% of atropine is eliminated unchanged in the urine. There is no correlation between plasma levels and maximal pharmacological effects after intravenous injections (Adams et al., 1982); therefore, the dose of atropine cannot be titrated by means of plasma concentration. For practical purposes, however, one should consider that the effects of intravenous atropine begin within 3–4 min and are maximal about 12–16 min after injection. Atropine is less effective in blocking certain muscarinic effects (for example, effects on the gastrointestinal and urinary tracts) than others effects (for example, effects on the heart and the salivary glands). Atropine has no effect on nicotinic symptoms, and central muscarinic effects may be undetectable, perhaps reflecting the difficulty of penetration of atropine into the CNS, which can be achieved only by large doses (Taylor, 1996a).

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Caution should be exercised in the use of atropine in hypoxic patients because it may cause ventricular fibrillation due to the increased myocardial oxygen demand brought about by the increased heart rate produced by atropine (Massumi et al., 1972). Therefore, in severe cases of OP poisoning, anoxia should be corrected before atropine is administered (Durham and Hayes, 1962). However, when arterial oxygen has been normalized, there is no reason to avoid the use of atropine because of the suggested risk of ventricular fibrillation (Kecik et al., 1993). Atropine is preferably given intravenously, although the intramuscular route is also effective. Satisfactory absorption also can be achieved by inhalation, and the pharmacokinetic characteristics are similar to those that occur after intramuscular injection (Harrison et al., 1986). However, atropine administration by inhalation has not been tested in cases of OP poisoning. Although several dosage regimens have been proposed and some caution was suggested in the dosage of atropine (de Kort et al., 1988), the best clinical approach is to administer doses of atropine large enough to achieve clinical evidence of atropinization, that is, flushing, dry mouth, changes in pupil size, bronchodilation and increased heart rate. If such signs are undetected, the dose of atropine is assumed to be insufficient and it must be increased (Barr, 1966). A mild degree of atropinization should be maintained for at least 48 h and withdrawal of atropine should be very carefully monitored because relapse can occur, particularly when OPs are stored in fat (see Section 72.1.3.1). In case of relapse, atropinization should be immediately reestablished. In patients with mild cholinergic signs, it is appropriate to start with a test dose of atropine (1 mg in the adult and 0.01 mg kg1 in children, intravenously). If signs of atropinization occur rapidly, severe poisoning is unlikely, although observation of the patient for at least 24 h is mandatory. In moderately to severely poisoned adult patients, 2–5 mg of atropine should be given intravenously and repeated every 10–20 min (0.02–0.05 mg kg1 in children at the same intervals). Continuous intravenous infusion may be required in severe cases. Because patients poisoned with OPs are tolerant to the effects of atropine, quite large doses of the drug have been used in cases of severe and prolonged poisoning (Golsousidis and Kokkas, 1985; Lotti et al., 1986). In a case of dimethoate poisoning, 30 g of atropine were given over 35 days with maximum daily dosage of 3.5 g (Le Blanc et al., 1986). Indicated dosages of atropine in accord with the severity of the clinical picture are summarized in Table 72.6. Overdosage with atropine is rarely serious in OP poisoned patients. On the contrary, patients frequently die because of insufficient atropine. When massive tachycardia is produced by atropine, it may be corrected by propanolol (Valero and Golan, 1967), thus avoiding the need to reduce the amount of atropine. It has also been suggested that using a combination of atropine and glycopyrrolate might offer an advantage over atropine

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Table 72.6  Indicative Dosage of Atropine in OP Poisoning According to Severitya Poisoning

Adults (mg)

Children (mg kg1 i.v.)

Mild

1.0

0.01

Moderate

2.0–5.0

0.02–0.05

Severe

20 h1

0.2 h1

(infusion)

(infusion)

a

See Table 72.5 for grading of poisoning severity.

alone, inasmuch as tachycardia could be avoided and adequate antimuscarinic effects still could be provided (Tracey and Gallagher, 1990), and that glucopyrrolate may better alleviate some signs of cholinergic overstimulation (Choi et al., 1998). Oximes  Oximes are nucleophilic chemicals that remove the phosphoryl group from the inhibited enzyme, thus restoring the catalytic site of AChE and its function (Bismuth et al., 1992). However, this chemical reaction occurs only when the phosphorylated AChE has not undergone the intramolecular rearrangement known as aging (Fig. 72.3 and Table 72.3; Aldridge and Reiner, 1972; Holmstedt, 1959; Taylor, 1996a). Whereas this reaction is fast (usually within a few minutes), oximes should be available at the synaptic cleft as long as there is newly inhibited AChE. Therefore, oximes should be administered over the first several hours after poisoning, although treatment may be prolonged in cases of massive poisoning and in poisonings by OPs with slow pharmacokinetics. Several oximes have been synthesized and are available (Bismuth et al., 1992; Dawson, 1994): pralidoxime (2-pyridine aldoxime or 2-PAM) is the most commonly used. Oximes currently in use are pralidoxime chloride (Protopam) in the United States, pralidoxime methylsulfate (Contrathion) in France and Italy, obidoxime chloride (Toxogonin) in Germany and Sweden, and pralidoxime methanesulfonate (P2S) in the United Kingdom. The availability of these oximes, including others such pralidoxime iodide, varies according to national pharmacopeas. Other oximes have been designed for the treatment of nerve gases, but they are not available for civilian uses (Kušic´ et al., 1985). The pharmacokinetics of oximes has been studied mostly in normal volunteers and a few studies have compared these results with those observed in poisoned patients. In studies that involved 15 volunteers given pralidoxime (single intravenous doses 2.5, 5.0, 7.5, and 10.0 mg), the plasma half-life was about 1.3 h and the apparent volume of distribution was 0.8 liter kg1 (Sidell and Groff, 1971; Sidell et al., 1972). Pharmacokinetic parameters have been derived after intravenous injection

of obidoxime in five healthy volunteers. The half-life (mean  SD) was 1.2  0.16 h and the volume of distribution (steady state) was 0.17 liter kg1 (Sidell et al., 1972). Pharmacokinetic data are also available for P2S (Holland and Parkes, 1976; Sundwall, 1960, 1961). A short infusion regimen of pralidoxime chloride was compared with administration of a loading dose followed by continuous infusion in healthy volunteers (Medicis et al., 1996). Plasma levels above 4 mg liter1 were maintained with the latter regime for twice as long as with the former (257 vs. 118 min). The pharmacokinetics of pralidoxime chloride was compared in nine healthy volunteers and six severely poisoned patients (Jovanovic´, 1989). Compounds involved in the poisonings were malathion, quinalfos, and dimethoate, and pharmacokinetic data were obtained after a single intramuscular dose of 1 g 2-PAM. The pharmacokinetic parameters (means  SD) in volunteers were different from those previously reported in the literature: plasma half-lives were 148  65 min, volume of distribution was 2.7 liter kg1, and total body clearance was 9  4ml min1 kg1. Nevertheless, when these parameters were compared with those in poisoned patients, significant differences were found. Thus, mean plasma concentrations of 2-PAM were one and a half times higher in patients than in volunteers at each time point and remained above 4 mg liter1 for 239 and 137 min, respectively. Elimination of oxime was greatly reduced in poisoned patients as compared with controls. Different pharmacokinetics of pralidoxime chloride also were reported in a series of children poisoned to a different degrees of severity with parathion, dichlorvos, diazinon, chlorfenvinfos, dicrotofos, and other OPs (Schexnayder et al., 1998). Continuous intravenous infusion of 10–20 mg kg1 h1 pralidoxime following a loading dose of 15–50 mgk g1 gave the following results: mean (SD) steady state plasma concentration of pralidoxime was 22.2 mg liter1  12.3 (range 6.9–47.4 mg liter1), mean (SD) plasma half-life was 3.6  0.8 h (range 2.4–5.3 h), mean clearance (SD) was 0.88  0.55 liter kg1 h1 (range 0.28–2.2 literkg1 h1), and mean volumeof distribution (SD)was5.5  4 liter kg1 (range 1.7–13.8 liter kg1). The large clearance and the high variability of volume of distribution found in these children should be noted. These differences of 2-PAM pharmacokinetics as compared with that in normal subjects are likely to be due to changes in hemodynamics during OP poisoning. A study that involved nine patients poisoned with various OPs (dimethoate, ethyl parathion, methyl parathion, and bromophos) indicated that pharmacokinetic data after intravenous pralidoxime methylsulfate were similar to those of pralidoxime chloride (Willems et al., 1992). Thus, after a loading dose of 4.42 mg kg1 pralidoxime methylsulfate, followed by a maintenance dose of 2.14 mg kg1 h1, plasma levels in these patients ranged from 2.12 to 9 mg liter1. Calculated pharmacokinetic data

Chapter | 72  Clinical Toxicology of Anticholinesterase Agents in Humans

(means  SD) were total body clearance of 0.57  0.27 liter kg1 h1, elimination half-life of 3.44  0.9 h, and volume of distribution of 2.77  1.45 liter kg1. Pharmacokinetic differences between healthy volunteers (see previous discussion) and poisoned patients also have been reported for obidoxime. Pharmacokinetic parameters were calculated under a steady state condition in the case of a patient severely poisoned with methamidophos and complicated by renal failure. This patient was given 4 mg kg1 obidoxime intravenously over 30 min every 6 h (Bentur et al., 1993). The obidoxime half-life was 6.9 h, the volume of distribution was 0.845 liter kg1, and the total body clearance 85.4 ml min1. Obidoxime plasma concentrations ranged from a preinfusion value of 5.6 g ml1 to 20.8 g ml1 during the infusion. These values are comparable to the reported values in healthy volunteers. On the contrary, the plasma half-life was longer, the clearance was lower, and the volume of distribution was larger. The most likely explanation for such differences is the renal insufficiency that affected this patient. In conclusion, pharmacokinetic parameters should be assessed cautiously because of changes in hemodynamics, particularly because of the reduction of renal blood flow produced by severe OP intoxication (Greener al., 1985; Jovanovi, 1989). Each oxime has different reactivation power on a given phosphorylated AChE and different phosphoryl residues attached to AChE are not equally susceptible to the same oxime (Aldridge and Reiner, 1972; Durham and Hayes, 1962; Kassa and Cabal, 1999; Worek et al., 1996, 1998a, b, 1999a). Several authors reported limited or no efficacy of oximes in the treatment of OP poisoning (Besser et al., 1995; Bismuth et al., 1992; de Silva et al., 1992; Erdmann et al., 1966; Singh et al., 1995; Tafuri and Roberts, 1987; Willems et al., 1993). Poisonings thought to be resistant to reactivation therapy involved chrotoxyphos, demeton, dimethoate, dimefox, methyl-phenkapton, shradan, prothoate, and triamiphos (summarized in Bismuth et al., 1992). Whereas these compounds form phosphoryl–AChE complexes that are identical to those formed by other OPs for which oximes have been found effective (Worek et al., 1999a), it seems to be inappropriate to consider resistance as the cause of lack of effects; other reasons should be sought, including the dose of oxime and duration of treatment, which are often insufficient (Johnson et al., 1992; Willems et al., 1993), the pharmacokinetics of the oxime, which may be quite variable depending on the clinical conditions of the patient, and the rates of oxime-induced reactivation of AChE (Worek et al., 1996). In this respect, the plasma concentration of the OP is relevant. In cases of ethyl and methyl parathion poisonings, oximes were shown to be ineffective as long as OP concentrations remained above 30 g liter1 (Willems et al., 1993; also see subsequent text). In one study, 21 severe to moderately poisoned patients treated with atropine alone were compared with 24 patients treated with atropine plus pralidoxime (de Silva et al., 1992).

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Patients were poisoned by a variety of OPs, including malathion, methamidophos, fenthion, dimethoate, phoxim, phentoate, and trichlorfon. Because the clinical outcomes were similar in the two groups, the authors cast doubt on the necessity of cholinesterase reactivators for the treatment of acute OP poisoning. This report was criticized because of the high mortality in both groups and because low doses of pralidoxime were used (Johnson et al., 1992). Moreover, given the specific characteristics of the phosphoryl residue attached to the enzyme in the rates of aging and of reactivatibility of AChE, the lack of measurement of plasma concentrations of OPs, and the different dosages of atropine, the study and the control groups are not comparable. Finally, there is no precise endpoint to assess the efficacy for a given oxime treatment when key variables that influence the clinical outcome are involved, such as mechanical ventilation. In conclusion, data on the inefficacy of oximes are not convincing and this potentially lifesaving therapy cannot be dismissed. On the contrary, such a course is highly recommended in any case of severe OP poisoning and treatment should continue as long as there is circulating OP. However, because OP plasma concentrations cannot be obtained easily, some empirical approaches have been suggested to assess the need for oxime treatment. For instance, measurements of RBC AChE before and after a bolus of oximes, or in vitro reactivatibility of AChE from RBC sampled from the patient may indicate whether newly inhibited AChE can be reactivated (Lotti, 1995). Other methods have been described based on in vitro inhibition of cholinesterases by the plasma of the patient (Dawson et al., 1997; Mahieu et al., 1982). Nevertheless, because reactivation of blood enzymes may not strictly reflect that in the synapses and because even minimal reactivation in the nervous system is likely to be beneficial, these approaches and their results should be regarded as a guide and not as a rule. Dosing regimes for various oximes that depend on the severity of poisoning have been suggested. The following regimes are recommended by the manufacturers. Pralidoxime chloride: Start with 1 g intravenously, followed by another 1 g after 15–30 min if no improvement. If still no improvement, start an infusion of 0.5 g h1. Slow intravenous administration is preferable, but intramuscular injection also is an option. In healthy volunteers, the pharmacokinetics of pralidoxime chloride was similar with either route of administration (Sidell and Groff, 1971). ● Pralidoxime methylsulfate: Start with 400 mg, followed by 200 mg after 0.5, 4, 6, and 12 h. In severe cases, start with 500 mg, repeat the same dose after 30 min, and then give 200 mg in repeated doses up to 2 g in 24 h. Continuous infusions also may be used (Willems et al., 1993), up to 500 mg/h in cases of slowly disposed OPs (Tush and Amstead, 1997). ●

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Obidoxime chloride: Start with 250 mg and repeat the same dose within 2 h, or 3–6 mg kg1 once or twice after poisoning. 1 ● In children, a loading dose of 25–50 mg kg pralidoxime chloride is recommended followed by a continuous infusion of 10–20 mg kg1 h1. All these recommended dosage schedules are aimed at achieving a plasma oxime concentration of 4 mg liter1, which was shown to be effective when using pralidoxime methanesulfonate (Sundwall, 1961). This concentration subsequently has been used a target reference value for all oximes. Nevertheless, this reference plasma level cannot be generalized because the molar concentrations of 2-PAM, obidoxime, and other oximes are different. Moreover, when the effects of various salts are compared, their different water solubilities should be taken into account because the amount of free base may vary considerably in the administered dose (Durham and Hayes, 1962). In some patients with ethyl and methyl parathion poisoning, enzyme reactivation could be obtained with pralidoxime methyl sulfate concentrations as low as 2.88 mg liter1, whereas in other patients, oxime concentrations as high as 14.6 mg liter1 had no effect. In such cases, the therapeutic effect of the oxime depends on the plasma concentration of ethyl and methyl parathion (Willems et al., 1993). Insufficient oxime therapy also has been considered as a possible cause of the intermediate syndrome (Benson et al., 1992). However, such a hypothesis does not explain why the intermediate syndrome is not influenced by atropine or why this condition selectivity affects certain neuromuscular junctions (see Section 72.2.3.1). Treatment with oximes is not reccommended in carbamate poisoning and in moderate poisoning by dimethylphosphate OPs because the rates of spontanous reactivation of dimethoxyphosphorylated and carbamylated AChE are fast. Moreover, with certain carbamates, more toxic complexes with oximes may be formed (Sterri et al., 1979). Pralidoxime is thought to be effective only on peripherally inhibited AChE because the quaternary nitrogen atom does not allow the drug to cross the blood–brain barrier. However, clinical observation in humans and some experimental data on animals point to the contrary (Lotti and Becker, 1982a; Namba et al., 1971). A possible explanation is an alteration of the blood–brain barrier brought about by OP poisoning. The toxicity of oximes was studied in human volunteers and iatrogenic effects were reported after treatment in some cases of OP poisoned patients (Marrs, 1991). In a study on volunteers, the acute and chronic toxicities of 2-PAM, P2S and 1, 1’-trimethylenebis(4-formylpyridinium chloride) (TMB4C12) were compared. TMB4C12 and, to a lesser extent, P2S (doses higher than 2.5 g a day) were found to cause marked gastrointestinal disturbances. Minor reversible cardiovascular effects also were noted (Calesnick et al., 1967). ●

Hayes’ Handbook of Pesticide Toxicology

In a patient severely poisoned with coumaphos, sudden cardiac arrest was observed 2 min after beginning an infusion of pralidoxime methyliodide, initiated about 4 h after poisoning. Trifluperazine and chloropromazine were also found in the blood of the patient. After restoring cardiac activity oxime infusion was started again but asystole recurred (Scott, 1986). The presence of phenothiazines might have potentiated the OP toxicity. However, this effect was shown in humans only in a single case report (Arterberry et al., 1962) and results from research on experimental animals are conflicting (Fernández et al., 1975; Michaleck and Stavinoha, 1978). Cardiac arrest was probably coincidental. Another case of worsening symptomatology coincident with the beginning of 2-PAM treatment was reported, although there were some notable differences (Good et al., 1993). A 51-year-old man presented at the hospital with limited signs consistent with moderate acute OP toxicity after several weeks of exposure to phosmet. RBC AChE was normal. Within several minutes of 2-PAM infusion, the patient had systemic and ventilatory weakness that required intubation. Electrophysiology revealed a subacute postsynaptic neuromuscular syndrome associated with some CNS dysfunctions that lasted for several weeks. One suggestive explanation offered by authors is that desensitization cholinergic receptors was produced by prolonged excess cholinergic stimulation and calcium influxes damaged the neuromuscular junction. The recommendation was made not to use oximes in the presence of postsynaptic dysfunction because of the possible direct effects of oximes on the receptor itself (Alkondon and Albuquerque, 1989). It remains to be explained why this patient had normal RBC AChE at the onset and during the entire course of the illness. There is some indication that further inhibition of AChE might occur during reactivation with 2-PAM (de Jong and Ceulen, 1978). Two phosphoryl oximes that formed during reactivation of the ethoxy methylphosphonyl–AChE conjugate by two oximes (LüH6 and TMB4) have been detected; they have not been detected during the reactivation of diethylphosphoryl–AChE conjugates. These phosphoryl oximes were found to be potent inhibitors of AChE, although usually they are likely to be hydrolyzed by paraoxonase (Luo et al., 1999). The worsening of symptomatology after treatment of massive poisonings with certain oximes may be due to accumulation of phosphoryloximes that occurs fast enough to saturate paraoxonase activity. It has been suggested that these events might also explain the pathogenesis of intermediate syndrome (Luo et al., 1999), although intermediate syndrome has been observed in many patients poisoned with diethyl OPs (see Section 72.2.3). Mild biochemical signs of liver toxicity have been related to the use of oximes. The symptoms disappeared when treatment was discontinued and seemed more frequent with obidoxime (Balali-Mood and Shariat, 1998; de Kort et al., 1988).

Chapter | 72  Clinical Toxicology of Anticholinesterase Agents in Humans

Diazepam  Diazepam must be included in the treatment of acute OP poisoning in all but the mildest cases (WHO, 1986). Diazepam relieves anxiety in mild cases, and reduces muscle fasciculations and antagonizes convulsions in the more severe cases (Johnson and Vale, 1992; Namba et al., 1971; Willems and Belpaire, 1992), although OP-induced convulsions are usually reduced by large doses of atropine (Vale and Scott, 1974). Moreover, animal data indicate that benzodiazepines improve morbidity and mortality in OP poisoning (Boškovic´ et al., 1984). Doses of diazepam (10–20 mg) given subcutaneously or intravenously are recommended and may be repeated as needed (Minton and Murray, 1988). Rarely, other anticonvulsants such as phenytoin have been used successfully in OP poisoning cases (Sellström, 1992). Supportive Treatment  The cornerstone of supportive treatment for severe poisoning is artificial ventilation, which must be started at the first signs of respiratory insufficiency. In such cases, admission to intensive care facilities is mandatory. Supplemental oxygen may be required to correct hypoxemia, and adjustments of fluid intake and electrolyte balance should be made as necessary. In severely ill patients, it may be necessary to maintain cardiac and urinary output pharmacologically. Prophylaxis of infections and ad hoc treatment of cardiac arrhythmias are also necessary. Sequence of Treatment  The sequence of first aid maneuvers depends on the circumstances and on the severity of poisoning. In cases of poisoning in the field, where antidotes are rarely available, the patient should be rushed to the hospital: very cautious decontamination may be tried in the case of dermal exposure, but in the case of ingestion, vomiting should not be induced. After the patient is evaluated in the hospital, the sequence of treatment is dictated by the severity of the clinical picture. Suggestions are given in Table 72.7.

72.1.3.9  Late Complications Neurological  Although it is known that the recovery time for some effects exceeds, to a limited extent, the time

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to replace AChE (Bowers et al., 1964; Namba et al., 1971; Whorton and Obrinsky, 1983), these cholinergic signs and symptoms that last up to several weeks after peak effects will not be considered. Moreover, OPIDP, a well established toxicity of some OPs, will be discussed separately (see Section 72.3). Several neurologic, psychiatic, and neurobehavioral abnormalities have been observed in patients who suffered previous acute poisoned with OPs, although prospective studies are not available. These symptoms and signs recently were conceptualized as a syndrome called chronic OP-induced neuropsychiatric disorders (COPIND; Jamal, 1997), together with similar symptoms also observed after long-term low-level exposures. COPIND after acute poisoning has been labeled phenomenon 1, whereas that after long-term exposure has been labeled phenomenon 2. Similarities between the two phenomena are few and superficial, many findings are contradictory and inconsistent, and given the different types of exposures, there is no reason to believe that they belong to the same entity. Moreover, possible neurological, psychiatric, and behavioral effects either after acute or after low-level long-term exposures would be better appreciated if a distinction is maintained (see Section 72.4.1). Like any condition associated with prolonged hypoxia, severe OP poisoning obviously could lead to various persistent neurological disorders of the CNS. Therefore, in cases of late CNS disturbances, assessments of the severity of the poisoning and of time elapsed between onset of symptoms and the beginning of treatment are required to distinguish between primary and secondary effects of OPs. For instance, a patient severely poisoned with sarin during the terrorist attack in Tokyo presented 6 months later with retrograde amnesia (Hatta et al., 1996). Because the patient was hypoxic for several minutes, it is impossible to ascertain whether the cause of amnesia was a direct biochemical effect of sarin. An unusual syndrome was reported in a patient 10 weeks after discharge from the hospital following acute poisoning with dimethoate (Sahin et al., 1994). The patient presented with erythema edema and hyperesthesia in the

Table 72.7  Suggested Sequence of Therapeutic Approaches to Acute OP Poisoning According to Severitya Clinical conditions

What to do First

Second

Third

Fourth

Mild

Decontamination

Atropineb (bolus)

Diazepamc

Observation

Moderate

Atropineb

Decontamination

Diazepamc

Pralidoximed

Severe a

Artificial ventilation

Diazepam

Modified from Lotti (1991). See Table 72.5 for grading severity. See Table 72.6 for dose. c 10 mg s.c. d See dose in the text. b

c

b

Atropine (infusion)

Pralidoxime

Fifth

d

Observation Decontamination

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hands associated with pain and limited movements. Upper arm electrophysiological studies revealed bilateral neuropathy and a bone scan detected increased osteoblastic activity of the hand bones. A diagnosis was made of reflex sympathetic dystrophy, which was associated with the previous poisoning. This was probably a coincidental association because the report was isolated, dimethoate does not cause OPIDP, and when OPIDP does occurs it does not exclusively affect the upper limbs or the bones (see Section 72.3). After ingestion of bromophos, a patient presented with no signs of cholinergic overstimulation and the only evidence of exposure was a reduction of plasma cholinesterases to about 10% of normal values. Oxime was given, but no atropine (Michotte et al., 1989). The patient was under treatment with maprotiline for a recurrent unipolar depressive disorder. Five weeks later the patient developed a cerebellar ataxia that subsided after 5 weeks. This case was likely a casual association, perhaps due to uncontrolled dosing with maprotiline. EEG changes in industrial workers with past repeated accidental exposures to the warfare agent sarin have been reported (Duffy et al., 1979). These exposures were not quantified, occurred at least 1 year prior to examination, and caused symptoms as well as significant RBC AChE inhibition. However, it is not clear whether cases of frank poisoning occurred. Some individuals had up to six such episodes. A number of differences, derived from complex analysis of EEG spectra, were observed between 77 exposed workers and 38 controls from the same factory but not exposed to sarin. These changes include increased  activity, increased  and  slowing, decreased  activity, and increased amounts of rapid eye movement sleep. Most of these changes were detected in the temporal and occipital lobes. Some controversies exist concerning the value of computerized analysis of brain wave topography (American Electroencephalographic Society, 1987; Duffy et al., 1986; Oken and Chiappa, 1986). In a commentary on these and similar results observed in animals (Burchfiel et al., 1976) the toxicological significance of these findings was questioned (Duffy and Burchfiel, 1980). The EEGs of 100 individuals with previous acute OP poisoning (one or more episodes occurring from 3 months up to 25 years before the survey) were compared with those of matched controls (Savage et al., 1988). Several OPs were involved in the poisonings, including methylparathion, parathion, malathion, disulfoton, mevinphos, dicrotophos, TEPP, dioxathion, DEF, and phorate. Poisoned cases had slightly more abnormal EEGs, but results were not significantly different between the matched cases and the control cohort. Although the authors stated that poisoning documentation was screened for completeness, some information was missing, such as the clinical severity of poisoning, the toxicological evidence of poisoning, and the nature of intercurrent diseases. For instance, one exclusion criterion was head trauma with period of unconsciousness

Hayes’ Handbook of Pesticide Toxicology

totaling more than 15 min, but the clinical conditions of cases with unconsciousness less than that were unreported. A retrospective study examined the vibrotactile thresholds in three groups of subjects: (1) previously poisoned with a variety of OPs (15 subjects), (2) poisoned with methamidophos (21 subjects), and (3) a matched control (35 subjects; McConnell et al., 1994). The results indicated that over one-fourth of the subjects previously poisoned with methamidophos, known to cause OPIDP (Senanayake and Johnson, 1982), had higher vibrotactile thresholds, but similar though less pronounced effects were seen in subjects poisoned with other OPs not known to cause OPIDP. The authors concluded that classical OPIDP is only the worst disease caused by methamidophos in a spectrum of peripheral nervous system impairments that represent the sequelae of poisoning. However, toxicological and clinical assessment of the poisoning episodes were not reported, and the elevated vibrotactile threshold was not symmetrical and also detected in the fingers. In mild toxic axonopathies, lesions are confined to the lower limbs and are characteristically symmetrical; hence, the described findings are not consistent with a toxic neuropathy. Moreover, subjects were examined 1–3 years after the acute poisoning and a toxic peripheral neuropathy is likely to recover over this period of time if exposure ceases. In a retrospective study (Steenland et al., 1994), 83 subjects exposed to a variety of OPs (among which only chlorpyrifos, which accounted for 10 cases, is known to cause OPIDP), who had one or more symptoms compatible with poisoning and documented inhibition of either RBC AChE or plasma BuChE (more than 20% of baseline or below normal range), showed significant alterations of vibrotactile sensitivity of fingers and toes compared with a control group (90 subjects). It is not clear when testing was performed, although it appears that it was done several years later. Because actual electrophysiological and vibration sensitivity data were not reported, and the clinical and toxicological data are not comprehensive, it is difficult to assess the biological significance of such changes. Moreover, as stated before, involvement of the arms is not expected in OPIDP unless it is extremely severe and a toxic peripheral neuropathy is expected to have recovered years after cessation of exposure. A similar retrospective study was reported by the same group of investigators (Ames et al., 1995). In 45 asymptomatic subjects who had a history of cholinesterase inhibition short of frank poisoning (RBC  70% of baseline or plasma cholinesterase 60% of baseline), some electrophysiological parameters were measured and no differences were found between cases and 90 controls. The OPs involved were not identified. The difference between cases in this study and those of the preceding study seems to be due to the presence of at least one symptom of cholinergic overstimulation in the subjects of the former study group, whereas the latter group had none. Cholinesterase inhibitions were overlapping

Chapter | 72  Clinical Toxicology of Anticholinesterase Agents in Humans

in the two studies. The authors concluded that preventing acute organophosphate poisoning also prevents neurological sequelae. Knowing the limitations of measurements of cholinesterase inhibition, of reporting symptoms common to both OP toxicity and to a variety of other conditions, and of electrophysiological studies on the upper limbs, and in the absence of detailed clinical data in one of the studies, it is difficult to compare the two studies or to agree with the authors’ conclusions about the existence of chronic sequelae of acute OP poisoning, which would not occur unless substantial AChE inhibition had occurred. In a study carried out in 1983–1984 and published in 1996, volunteers exposed to sarin concentrations that caused 30–40% RBC AChE inhibition showed mild electrophysiological changes up to 15 months after a single exposure. The small number of subjects, the high variability of changes, and the fact that only three out of eight subjects displayed persistent changes make this study difficult to interprete (Baker and Sedgwick, 1996). In a cross-sectional study on 164 pesticide workers, a correlation was found between past OP poisoning and increased incidence of symptoms such as dizziness, sleepiness, and headache (London et al., 1998; see also Section 72.4.1.2). No evidence of correlation with past poisoning was found when vibration sense and tremor were evaluated and found to be unaltered. Moreover, clinical details of OP acute poisoning were not given. Psychiatric  Follow up studies based on interview, physical examination, and blood chemistry on long-term sequelae of acute OP poisoning revealed no significant serious neuropsychiatric effect in a group of 114 individuals, 6 of whom had severe poisoning and the others mild to moderate poisoning (Tabershaw and Cooper, 1966). An array of different symptoms reported by subjects was not considered to be associated with the poisoning. However, the authors conceded that their study would not reveal minor after effects or those of low incidence. Preexisting psychiatric symptomatology has been reported to have worsened over 2 years after OP poisoning and whenever small further exposures to OPs occurred (Rosenthal and Cameron, 1991). Details of the poisoning of this patient were not given. Neurobehavioral  In the study described previously, combined clinical and neuropsychological evaluations were used to detect changes in the cognitive functions in a group of 100 subjects with previous acute OP poisonings compared with a matched control group (Savage et al., 1988). In this study, however, the limitations already outlined raise the question whether these changes represent a consequence of brain hypoxia or of other intercurrent factors, given the very large variability in the time elapsed from poisoning to assessment. Most differences between the two groups, detected on a number of tests, were within normal variability. Certainly other factors such as educational

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differences might account for differences in comprehension, arithmetic, vocabulary, etc. Moreover, toxicological analysis showed that blood levels of organochlorine pesticides in the study group were about twice those of the controls. Whereas statistical analysis failed to show any association between such blood levels and the results of neuropsychological tests, the authors ruled out organochlorine as the causative agent of such impairments. However, from a toxicological viewpoint, organochlorine exposure might have been more relevant, given the pharmacokinetic differences between these pesticides and OPs, especially because organochlorine is more easily stored in the body. In a retrospective cohort study (Rosenstock et al., 1991), a group of 36 subjects previously poisoned with OPs were tested on average about 2 years after the episode of poisoning and compared with a matched control group. The poisoned group did much worse than the control group on several neuropsychological tests (visual and verbal attention, visual memory, visuomotor activities, and dexterity). The type of OPs involved and the severity of poisoning were not reported, and the design and statistical significance of the study were criticized (Schuman and Wagner, 1991). Moreover, subjects in the control group (a close friend or sibling in the same community who had never been treated for pesticide poisoning) were also occupationally exposed to OPs. Given the endpoints used (for instance, visuomotor performance) and the lack of follow-up studies, it is possible that the neuropsychological deficits were a cause rather than a consequence of OP poisoning. A neuropsychological test battery was administered to 21 migrant farm workers who had been acutely exposed to phos-drin and other pesticides (lannate and maneb) and to matched controls (Reidy et al., 1992). Two acute exposures occurred 3 years apart and subjects were examined 2 years after the second exposure on the occasion of a worker’s litigation. The exposed group was significantly more impaired than controls on tests of psychomotor speed, dexterity, and visuospatial memory. Although symptomatic, RBC AChE and plasma BuChE were normal on both occasions. Therefore, if related these changes were to pesticide exposures, they cannot be a consequence of OP poisoning, but perhaps of the other involved pesticides. In the previously quoted study (Steenland et al., 1994), several behavioral parameters were tested, but only sustained activity was found to be worse in the case group than in the controls. In another previously mentioned study (Ames et al., 1995), a number of neurobehavioral tests were performed on subjects who had a history of cholinesterase inhibition “short of frank poisoning.” Only one test (serial digit performance) was statistically significant, but it was opposite to the hypothesized direction. In conclusion, there is little evidence to support the notion that acute OP poisoning may result in late permanent toxic effects other than OPIDP if hypoxia and/or severe uncontrolled convulsions did not occur or did not last for

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sufficient time. Similar conclusions have been reached by others (Ray, 1998a, b). Moreover, ranking all these effects under one syndrome (COPIND phenomenon 1) is inappropriate and may be misleading.

72.2  The intermediate syndrome The intermediate syndrome is characterized by weakness of respiratory, neck, and proximal limb muscles. It is not a direct effect of AChE inhibition and appears several hours after the beginning of signs and symptoms of severe cholinergic over-stimulation. It is caused by a variety of OPs and seems to be related to postsynaptic effects. This form of OP toxicity was first conceptualized by Senanayake and Karalliedde (1987), although the first accurate description of the syndrome was given by (Wadia et al., 1974), who categorized the neurological manifestations of acute OP poisoning into two groups. Type 1 signs are the classical signs of cholinergic overstimulation, whereas type 2 signs, which appear later and while undergoing atropine treatment, are characterized by proximal weakness and cranial nerve palsies. New case reports and prospective studies or cases derived from retrospective analysis of medical records have been reported.

72.2.1  Etiology Intermediate syndrome seems to occur in 20–50% of acute OP poisoning cases (Sedgwick and Senanayake, 1997). It has been observed after exposures to several OPs, including fenthion (De Wilde et al., 1991; Karademir et al., 1990; Senanayake and Karalliedde, 1987), omethoate (He et al., 1998), dimethoate (De Bleecker et al., 1993; He et al., 1998; Senanayake and Karalliedde, 1987), methamidophos, monocrotophos (Senanayake and Karalliedde, 1987), diazinon (Wadia et al., 1974), demeton S-methylsulfone (Besser et al., 1989a), trichlorfon (Karademir et al., 1990), parathion (De Bleecker et al., 1993; He et al., 1998), methy­lparathion (De Bleecker et al., 1993) dichlorvos (He et al., 1998), phosmet (Good et al., 1993), and malathion (Gadoth and Fisher, 1978), and to various mixtures of OPs (De Bleecker et al., 1993; He et al., 1998).

72.2.2  Pathogenesis The mechanism by which the intermediate syndrome develops is unknown. The first characterization of the syndrome suggested a postsynaptic effect based on electromyographic evidence of fade on tetanic stimulation, absence of fade on low-frequency stimulation, and absence of posttetanic facilitation (Senanayake and Karalliedde, 1987). This concept was further supported by morphological and electrophysiological studies, both in humans

Hayes’ Handbook of Pesticide Toxicology

(De Bleecker et al., 1992b, 1993; Sedgwick and Senanayake, 1997; Singh et al., 1998a, b) and animals (Engel et al., 1973), suggesting that muscle weakness may result from cholinergic receptor desensitization due to prolonged cholinergic stimulation. Therefore, the hypothesis was put forward that the pathophysiology of intermediate syndrome is the result of a time-confined phenomenon that includes both changes in the postsynaptic structures by desensitization and restoring the ratio of acetylcholine to AChE (De Wilde et al., 1991). This process may explain the observation of an unusual case of respiratory failure precipitated by 2-PAM in a patient thought to have had prolonged cholinergic overstimulation by phosmet (Good et al., 1993). Neuromuscular block may have been increased because of a sudden reduction of acetylcholine levels that had caused the postsynaptic dysfunction. Similarly, a patient poisoned with oxydemeton-S-methyl was comatose shortly after poisoning, but responded to noxious stimuli. However, such response decreased suddenly 42 h after intoxication and the electrophysiological investigation performed 66 h after poisoning detected severe neuromuscular dysfunction characterized by the decrement phenomenon after repetitive nerve stimulation. At this point in time, obidoxime was given and the neuromuscular block worsened (Besser et al., 1995). With this interpretation, the hypothesis that the intermediate syndrome may result from excessive and persistent acetylcholine levels due to insufficient oxime therapy during the early acute phase of cholinergic toxicity might have some validity (Gadoth and Fisher, 1978; Benson et al., 1992). Another hypothesis relates the development of intermediate syndrome to the formation of oxime–phosphoryl complexes, which greatly inhibit AChE and not efficiently cleaved by paraoxonase (Luo et al., 1999; see Section 72.1.3.8). Muscle necrosis, the human equivalent of a well known effect observed in experimental animals (Dettbarn, 1992), was also thought to be a possible cause of intermediate syndrome. Myopathy was described by de de Reuck and Willems (1975), Ahlgren et al. (1979), de Reuck et al. (1979), and Wecker et al. (1985; see Section 72.1.3.6). However, the clinical evidence of muscle necrosis was not consistent in a 21 case series because only about one-half of the patients had elevated CPK and LDH (He et al., 1998). Moreover, the histopathological lesions were too limited (De Bleecker et al., 1993) to support the notion that intermediate syndrome is due to muscle necrosis. In conclusion, in support of the explanation that intermediate syndrome is a consequence of desensitization of nicotinic receptors is the observation that several OPs, quite different chemically and toxicologically, did produce intermediate syndrome and that all patients had a prolonged AChE inhibition and consequent high levels of acetylcholine. However, the reasons for the selectivity for some nicotinic receptors only, as shown by the clinical features of intermediate syndrome, remain unexplained.

Chapter | 72  Clinical Toxicology of Anticholinesterase Agents in Humans

72.2.3  Clinical Manifestations 72.2.3.1  Clinical Signs and Course The intermediate syndrome develops during recovery from cholinergic manifestations, one to several days after the poisoning. Distinction should be made between this syndrome and the recurrence of cholinergic toxicity, which may occur with OPs that display prolonged disposal (Davies et al., 1975; Ecobichon et al., 1977; Gadoth and Fisher, 1978; Molphy and Rathus, 1964; Perron and Johnson, 1969; see Section 72.1.3.1). In some cases, the sudden onset of intermediate syndrome occurs in patients when they are completely recovered from the initial cholinergic crisis (Senanayake and Karalliedde, 1987; Wadia et al., 1974), whereas in others, it is concurrent with muscarinic signs of toxicity or with superimposed muscarinic relapses (De Bleecker et al., 1993). Concurrent and recurrent cholinergic signs are controlled by atropine, whereas those of the intermediate syndrome are not. A constant feature in all patients is a marked weakness of neck flexion and of proximal limb muscles. These patients are unable to raise their head off the pillow, to abduct their shoulders’ or to flex their hips. A sudden respiratory failure due to weakness of respiratory muscles also characteristically occurs in 70–100% of the cases, often drawing attention to the onset of the syndrome. Upper and lower limb reflexes are often reduced or absent, and in several patients there was evidence of weakness of muscles innervated by cranial nerves. One or more nerves may be involved, including the III, IV, VII, IX, X, and XI. There is no distinct pattern in the development of these signs. Mortality due to respiratory paralysis and complications ranges from 15 to 40%. The clinical course in surviving patients lasts up to 30–40 days. Regression of cranial nerve palsies appears first, followed by improvement of respiratory insufficiency and recovery of strength in the proximal limb muscles. Neck flexion is the last function to recover.

72.2.3.2  Electrophysiology Electrophysiological studies have been performed on a few patients with intermediate syndrome (De Bleecker et al., 1993; Sedgwick and Senanayake, 1997; Senanayake and Karalliedde, 1987; Singh et al., 2000; Wadia et al., 1987). Nerve conduction velocity was normal and distal latencies were either normal or slightly reduced. No signs of spontaneous activity, such as fibrillation potentials or positive sharp waves, were observed. Repetitive stimulation showed decrements of CMAP at low and/or intermediate frequencies in most patients (10–50 Hz), lasting for several days, but always disappearing before clinical normalization, indicating that the neuromuscular junctional dysfunction in the intermediate syndrome is likely postsynaptic. Electrophysiological changes that occur after repetitive stimulation have been detected both in poisoned patients

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with clinical evidence of intermediate syndrome and in patients with acute poisoning without subsequent development of intermediate syndrome (De Bleecker et al., 1993). In the latter group, these changes followed recovery from initial depolarization block due to acetylcholine excess (see Section 72.1.3.5). It is, therefore, not clear whether this postsynaptic block is always present in all cases of acute OP poisoning, beginning a few days from the initial cholinergic syndrome. In such a case, the reasons for the switch from subclinical to a clinically evident intermediate syndrome are unknown.

72.2.3.3  Pathology Muscle histopathology was performed in very few patients (De Bleecker et al., 1992a, b, 1993). Few and scattered necrotic muscle fibers were observed. Histochemical endplate AChE staining was variable in intensity. Ultrastructural examination showed neuromuscular junctions that contained numerous synaptic vesicles, swollen mitochondria, and synaptic clefts with well established basal lamina. Some endplates showed vesiculations and phagocytic lysosomal activity suggestive of degeneration. Some of the synaptic clefts were widened and filled with debris, junctional folds were simplified, and postsynaptic areas were denuded and degenerated.

72.2.3.4  Treatment Treatment is exclusively supportive because there is no specific treatment for the intermediate syndrome and atropine is not effective. Endotracheal intubation and mechanical ventilation are lifesaving.

72.3  Delayed polyneuropathy Organophosphate-induced delayed polyneuropathy is a rare toxic effect in humans, although some epidemics have occurred in the past such as the famous Ginger-Jake paralysis when thousands of patients were intoxicated with triorthocresyl phosphate (TOCP; Morgan, 1982; Inoue et al., 1988). Neuropathy is characterized by a symmetric, distal sensory–motor, central–peripheral axonopathy that affects the legs and, in the most severe cases, also the arms. OPIDP is mechanistically unrelated to cholinergic and intermediate syndromes, and, therefore, it is not necessarily associated with the anticholinesterase activity of OPs. In fact, several OPs (such as triarylphosphates) are devoid of this activity, but may cause OPIDP. A large body of experimental data indicate that this axonopathy is likely to be correlated with effects on a neural esterase known as neuropathy target esterase (NTE; Johnson, 1990; Lotti, 1992b). Clinical onset is delayed for up to 10–20 days after a single exposure and for an unspecified period after continuous exposures.

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72.3.1  Etiology Not all OPs are capable of causing OPIDP and in the case of OP insecticides currently in use, polyneuropathy develops exclusively after a severe episode of cholinergic toxicity. Because OPIDP is not a consequence of acute cholinergic toxicity, but just another toxic effect of OPs, in the case of insecticides it will develop at a much higher dose than that which causes cholinergic overstimulation. This is indirectly shown in Table 72.8, where sensitivity to various inhibitors of AChE (target of cholinergic toxicity) and of NTE (target of OPIDP) derived from human tissues are compared. OP insecticides that caused OPIDP in humans (Table 72.9) produced OPIDP only after cholinergic toxicity. Mipafox was never developed as an insecticide and caused OPIDP after mild cholinergic toxicity. Phenylsaligenin phosphate is the active metabolite of TOCP (not an insecticide), which caused several cases of OPIDP without cholinergic toxicity. Similar differences were seen with hen enzymes, which in turn correlate with the capability of a given OP to cause OPIDP, relative to that of causing death (Lotti and Johnson, 1978). Therefore, compounds with AChE I50/NTE I50 ratios  1 may cause OPIDP without cholinergic toxicity, whereas those with a ratio  1 will cause OPIDP only after cholinergic toxicity and appropriate antidotal treatment. OPIDP displays a characteristic age-related sensitivity in both experimental animals and humans. Children are resistant to OPIDP (Goldstein et al., 1988) and when they are affected, they recover much quicker than adults, usually within a few months (Senanayake, 1981). In addition to the case reports listed in Table 72.9, other reports can be found where development of OPIDP was associated with single or short-term exposures to certain OPs, although clinical and toxicological evidence was not convincing. EPN and leptophos cause OPIDP in hens and are NTE inhibitors, but only at doses that cause severe cholinergic toxicity (Johnson, 1975; Ohkawa et al., 1980). However, a report of OPIDP that involved several workers who had long-term exposure to EPN indicated little or no evidence of cholinergic overstimulation and most clinical details were missing. Moreover, during the release of EPN, these workers were also exposed to other chemicals derived from an explosion and fire in a manufacturing facility (Xintaras and Burg, 1980). An outbreak of neurological disorders occurred in a plant that manufactured leptophos (Xintaras et al., 1978), which is known to cause OPIDP in experimental animals (Hollingshaus et al., 1981). Three subjects had signs compatible with OPIDP at medical examination and six had symptoms in a retrospective study. Several subjects, however, had neurological signs unrelated to OPIDP and all were exposed to a variety of neurotoxic chemicals, including n-hexane. Another group of case reports suggested OPIDP development after exposures to omethoate, parathion, mecarbam,

Table 72.8  Comparative Sensitivities of Human AChE and NTE for Various Inhibitorsa Compoundb

AChE I50 (M)

NTE I50 (M)

AChE I50

Dichlorvos

0.95

16

0.06

Methamidophos (I isomer)

100

400

0.2

Chlorpyrifos-oxon

0.01

0.2

0.05

Mipafox

100

12

1

Phenylsaligenin phosphate

0.12

0.003

39

NTE I50

a Data from Lotti and Johnson (1978), Bertolazzi et al. (1991), and Capodicasa et al. (1991). b Direct acting OPs or metabolites (chlorpyrifos oxon from chlorpyrifos and phenylsaligenin phosphate from TOCP).

fenthion, and mevinphos. However, these pesticides are not NTE inhibitors and negative results have always been reported in the hen test (FAO/WHO, 1987, 1996, 1997; Johnson, 1975; Lotti, 1992b). A polyneuropathy compatible with OPIDP developed after a suicide attempt with omethoate (Curtes et al., 1981), but in a subsequent report of a man who died shortly after an acute omethoate poisoning, no NTE inhibition was detected in post mortem nerve tissues (Lotti et al., 1981). Parathion was associated with OPIDP after massive poisoning together with methanol (de Jager et al., 1981). Toxicological evidence of parathion in body fluids was missing and the clinical description does not support evidence of RBC AChE inhibition or methanol poisoning (Lotti and Becker, 1982b). Moreover, several cases of severe parathion poisoning resulted in no OPIDP (Namba et al., 1971). Mecarbam was reported to cause neuropathy, but nerve biopsy revealed segmental demyelination without axonal degeneration (Stamboulis et al., 1991), a morphological lesion not expected in OPIDP (see Section 72.3.3.3). A case of delayed neuropathy apparently developed 1 month after acute poisoning by fenthion, which was followed by intermediate syndrome (Karademir et al., 1990). It is not clear whether diagnosis was made on clinical grounds or exclusively on electromyography (EMG). Because the results of EMG studies were not reported, interpretation of this case is difficult. One further case of delayed polyneuropathy by fenthion was reported (Martinez-Chuecos et al., 1992), but no clinical details were given. Moreover, another patient poisoned with fenthion, and belonging to the same series, did not develop neuropathy. A female patient who attempted suicide with methylparathion, fell into a deep coma that lasted 4 weeks (Nisse et al., 1998). Electromyography was normal 3 weeks after

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Table 72.9  OP Insecticides that Cause Delayed Polyneuropathy in Humansa Compound

Number of cases

Circumstances

References

Chlorpyrifos

2

Suicide

Lotti et al. (1986); Martinez-Chuecos et al. (1992)

Dichlorvos

3

Suicide

Vasilescu and Florescu (1980); Wadia et al. (1985)

Isofenphosb

3

Suicide

Catz et al. (1988); Tracey and Gallagher (1990); Moretto and Lotti (1998)

Methamidophos

Several

Suicide/occupational

Senanayake and Johnson (1982); Moretto and Lotti (1998); McConnell et al. (1999)

Mipafox

2

Occupational

Bidstrup et al. (1953)

Trichlorfon

Several

Suicide

Hierons and Johnson (1978); Johnson (1981); Shiraishi et al. (1983); Niedziella et al. (1985); Csik et al. (1986)

Trichlornate

2

Suicide

Jedrzejowska et al. (1980); Willems (1981)

a

Modified from Lotti (2000). All cases displayed preceding cholinergic toxicity. One of these cases was a combined exposure with phoxim.

b

poisoning, whereas 1 week later it showed signs of mild distal sensory motor polyneuropathy. Diffuse myogenic alterations were also detected, but electrophysiological data were not reported. Because the neuropathy disappeared within 4 weeks, this was unlikely a case of OPIDP and was probably a consequence of prolonged coma. A case of severe poisoning by mevinphos was reported to have been complicated by polyneuropathy (Hsiao et al., 1996). No clinical or electrophysiological data were given. It is said that nerve conduction studies confirmed the neuropathy. In such a case, OPIDP would be unlikely, because conduction is usually, at the most, slightly affected in axonopafhy. No followup was reported. An isolated case of Guillan-Barré-like syndrome was described in a patient after exposure to merphos, a defoliant with little anticholinesterase activity (Fisher, 1977). Exposure was dermal and likely very low. Four days later the patient started complaining of upper and lower limb weakness; 14 days after exposure, he developed facial diplegia. EMG and clinical features were suggestive of Guillan-Barré syndrome. This is the only case in the literature of Guillan-Barré-like signs after acute exposure to OP, but the clinical and electrophysiological characteristics of OPIDP are quite different. In conclusion, combined clinical and experimental evidence allows firm conclusions on OPIDP only for a few chemicals (Table 72.9). Nevertheless, it should be pointed out that neuropathic impurities may be present in commercial formulations, which perhaps accounts for some of these cases (Johnson, 1984).

molecular changes occur within a few hours of exposure, but almost nothing is known about what happens between these events and the clinical, morphological, and electrophysiological onset of OPIDP 2–3 weeks later (Johnson, 1990; Lotti, 1992b). Limited evidence suggests that development of OPIDP in humans also involves inhibition of NTE. In two fatal cases of OP poisoning, NTE activity was measured post mortem in the peripheral nerve. As expected from animal studies, NTE was found to be inhibited in a case of chlorpyrifos poisoning (Osterloh et al., 1983), whereas it was not in a case of omethoate poisoning (Lotti et al., 1981). Lymphocytic NTE (see Section 72.3.3.2) was found to be inhibited soon after exposure in one case of poisoning with chlorpyrifos (Lotti et al., 1986) and in two cases with methamidophos that all developed OPIDP weeks later (Moretto and Lotti, 1998; McConnell et al., 1999). Lymphocytic NTE inhibition was also found in a patient poisoned with isofenphos who died on day 32; OPIDP might have developed after this time. In severe poisoning by compounds not known to cause OPIDP, lymphocytic NTE inhibition was not detected (Moretto and Lotti, 1998). After occupational exposures to DEF, substantial NTE inhibition in lymphocytes was measured, but found not to be associated with the development of OPIDP or with electrophysiological changes. In this case, lymphocytic NTE did not represent a good mirror of peripheral nerve NTE, probably because of the particular pharmacokinetics of this compound (Lotti et al., 1983).

72.3.3  Clinical Manifestations 72.3.2  Pathogenesis

72.3.3.1  General Features

The initial event in OPIDP is the inhibition of NTE, followed by aging of the phosphoryl NTE complex. These

Symptoms of OPIDP begin 2–3 weeks after single doses when, as in the case of insecticides, cholinergic symptoms

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have subsided (Lotti et al., 1984). The lag time between single or short-term exposure and the clinical onset of OPIDP depends on both the chemical involved and the dose (Bidstrup et al., 1953; Lotti et al., 1986; Senanayake and Johnson, 1982). OPs with slow pharmacokinetics may cause OPIDP after a prolonged period following exposure (up to 4 weeks), whereas higher doses of OPs that are powerful in causing OPIDP may shorten this period to about 10 days. Clinical features of OPIDP are usually fully expressed within a few days of the onset of symptoms and signs, and no progression has been observed in the absence of further exposure. After repeated exposures, such as those to nonanticholinesterase OPs, the onset of symptoms and their full development is more variable and less definible (Vasilescu, 1982). The usual initial complaint is cramping muscle pain in the legs (Susser and Stein, 1957), followed by distal numbness and paresthesia (Senanayake and Jeyaratnam, 1981; Vasilescu et al., 1984). Progressive leg weakness occurs, together with depression of tendon reflexes. Symptoms and signs may also appear in the arms and forearms following those in the legs, but always after severe exposures (Bidstrup et al., 1953; Moretto and Lotti, 1998; Senanayake and Jeyaratnam, 1981; Vasilescu, 1982; Vasilescu et al., 1984). Physical examination reveals distal symmetrical predominantly motor polyneuropathy, with wasting and flaccid weakness of distal limbs muscles, especially in the legs. Signs include a characteristic high-stepping gait associated with bilateral foot drop (Senanayake and Jeyaratnam, 1981). Severe OPIDP may result in quadriplegia with foot and wrist drop as well as mild pyramidal signs. In time, there is complete functional recovery if spinal cord axons have been spared by smaller doses (Senanayake, 1981); otherwise, pyramidal and other signs of central neurological involvement may become more evident. The degree of pyramidal involvement determines the prognosis for functional recovery, and spastic ataxia may represent a permanent outcome of severe OPIDP (Morgan and Penovich, 1978; Tosi et al., 1994; Vasilescu, 1982). Objective evidence of sensory loss is usually slight or absent. In one group of patients poisoned with methamidophos, some sensory symptoms, but no objective signs, were recorded (Senanayake and Johnson, 1982). In two patients who developed OPIDP after exposure to chlorpyrifos and isofenphos, slight sensory alterations were detected during both physical and electrophysiological examination (Moretto and Lotti, 1998). However, in a series of patients, cases were reported where purely sensory peripheral neuropathy was displayed after repeated low exposures to chlorpyrifos that caused some symptomatology, and no signs or mild signs of cholinergic overstimulation (Kaplan et al., 1993). This contrasts with the known toxicological characteristics of chlorpyrifos, which is a better inhibitor of AChE than NTE (Capodicasa et al., 1991; Richardson, 1995), and the clinical features observed in two cases of

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OPIDP induced by chlorpyrifos, where OPIDP was always preceded by severe cholinergic overstimulation (Lotti et al., 1986; Martinez-Chuecos et al., 1992). Whereas the exposure assessment in the Kaplan series was limited and based almost exclusively on medical history, interpretation of these discrepancies is difficult.

72.3.3.2  Laboratory Findings They are no specific changes in common laboratory tests, including chemical and morphological analysis of spinal fluid. Increased serum levels of immunoglobulin G autoantibodies to glial fibrillary acidic protein and to neurofilament 200 have been detected in a case of methamidophos poisoning after the development of OPIDP (McConnell et al., 1999), probably reflecting peripheral nerve damage. Lymphocytic NTE  NTE activity was found in lymphatic tissues in humans (Moretto and Lotti, 1988) and its characteristics in circulating lymphocytes led to the conclusion that the level of this blood enzyme is similar to that in the nervous system (Bertoncin et al., 1985). On this basis, suggestions were made to measure and use lymphocytic NTE like RBC AChE activity is used in the clinical setting and in the biomonitoring of occupational exposures (Lotti, 1987). NTE activity has also been detected in humans platelets (Maroni and Bleecker, 1986). Only on one occasion has the ratio between NTE inhibition in lymphocytes and peripheral nerves been measured, and it was found to be about 1 (Osterloh et al., 1983), although it is anticipated that it will not be always so, given the different pharmacokinetics of OPs. Inhibition of lymphocytic NTE soon after poisoning was predictive of OPIDP development when measured several days before the onset of OPIDP in two cases (McConnell et al., 1999; Moretto and Lotti, 1998). Given the relatively high turnover of blood lymphocytes and the usually rapid disappearance of OPs from the blood, measurement of lymphocytic NTE should be made in the early days after poisoning because the activity may be back to almost normal at the onset of OPIDP. However, because no treatment for OPIDP is available, the detection of lymphocytic NTE soon after poisoning has limited clinical value. Similarly, measurements of lymphocytic NTE to monitor occupational exposures to OP insecticides have no practical value because such exposures preferentially inhibit blood cholinesterases and, therefore, lymphocytic NTE rarely would be affected (see Section 72.3.1). Electrophysiology  Electrophysiological changes are usually detected concurrently with the onset of clinical symptoms and signs of OPIDP. When performed during the symptom-free period between the disappearance of cholinergic toxicity and the clinical onset of OPIDP, the electrophysiological examination is normal (Lotti et al., 1986). The electrophysiological picture accords well with the histopathological findings of distal axonopathy (Jedrzejowska et al., 1980; Lotti et al., 1986; Moretto and Lotti, 1998;

Chapter | 72  Clinical Toxicology of Anticholinesterase Agents in Humans

Vasilescu and Florescu, 1980; Vasilescu et al., 1984; see Section 72.3.3.3). The evaluation reveals partial denervation of affected muscles, with increased insertional activity, abnormal spontaneous activity (fibrillation potentials and positive sharp waves), and a reduced interference pattern; large polyphasic motor unit potentials also may be found after a few weeks. The compound muscle action potentials to supra-maximal stimulation of motor nerves are reduced in amplitude, and terminal motor latencies are delayed; maximal conduction velocity is usually normal or slightly reduced. Minimal electrophysiological abnormalities of sensory function are occasionally detected. About 1 year after poisoning, normalization of electrophysiological parameters parallels that of clinical signs unless the pyramidal tract is involved. In such a case, findings may resemble those of amyotrophic lateral sclerosis (Vasilescu, 1982).

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difficult is differential diagnosis when substantial exposures to nonanticholinesterase OPs that cause OPIDP is overlooked. Symmetrical leg involvement with additional involvement of upper limbs only in severe cases, lack of involvement of cranial nerves and the autonomic system, and electromyographic changes consistent with distal axonal neuropathy are all indicative of OPIDP. Medical history aimed at identification of possible sources of OP exposure remains, in these cases, the only way to etiologically attribute neuropathy.

72.3.3.5  Treatment There is no specific treatment for OPIDP Intensive programs of physical therapy are indicated to ameliorate muscle trophism during the recovery from peripheral nerve lesions. In later stages, if spasticity develops, GABA antagonists may be used.

72.3.3.3  Pathology The histopathology of OPIDP has rarely been described in humans (Aring, 1942; Jedrzejowska et al., 1980; Lotti et al., 1986; Vasilescu et al., 1984), although there are no major differences from what has been extensively observed in experimental animals (Abou-Donia and Lapadula, 1990; Cavanagh, 1973; Tanaka and Bursian, 1989). The central peripheral distribution of lesions is similar to that of toxic neuropathies of other origins. The vulnerability of nerve fibers is directly related to axonal length and diameter; large-diameter and long fibers are more susceptible than small and short ones. Spinal cord changes involve mainly the anterior horn cells and the pyramidal tracts. Lesions in the tract of Goll were less constant and no lesions were seen in the tract of Burdach (Aring, 1942). Sural nerve biopsies indicated axonal-type lesions with an even degree of involvement of myelinated fibers of different sizes and a lesser degree of involvement of unmyelinated fibers. Dark and swollen axoplasm due to the accumulation of axoplasmic organelles can be observed association with aspects of axonal degeneration. On teased fiber preparations, some ovoids arranged in linear rows were identified. Electron microscopy found myelin debris in the Schwannian profiles. Depending on the time of biopsy after poisoning, various stages of regeneration and remyelination can be observed. Segmental demyelination is not observed (Jedrzejowska et al., 1980; Lotti et al., 1986). These changes indicate a process that is a primary distal axonopathy with moderate, secondary, and distal loss of myelin.

72.3.3.4  Differential Diagnosis The unequivocal suggestion for diagnosis of OPIDP caused by insecticides is the presence of an episode of acute cholinergic toxicity in the recent past medical history. More

72.4  Long-term exposures Long-term exposures to OPs may cause cholinergic syndrome if both the size of repeated doses and the intervals between them overcome AChE resynthesis in the nervous tissues. In such a case, a buildup of AChE inhibition may occur and when threshold is reached, symptoms that are indistinguishable from those observed after single or shortterm exposures are produced. Therefore, only signs and symptoms unrelated to overt cholinergic toxicity will be considered in this section. Many reports on the effects of long-term exposures to OPs lack follow up studies, particularly after cessation of exposures. Therefore, most of the described effects may not be chronic (i.e., longstanding or irreversible) and probably reflect the effects of current exposures. Therefore, it is advisable not to talk about chronic effects, but rather of effects of low-level exposures, either during exposure or shortly afterward, and keep them distinct from effects detectable several months or years after cessation of exposure. Moreover, the major problem of these studies is often the insufficient assessment of exposures, that obviously hampers the interpretation of results.

72.4.1  Neurological, Psychiatric, and Behavioral Effects The large amount of literature that describes neurological, psychiatric, and behavioral effects has been reviewed in several articles (Brown and Brix, 1998; ECETOC, 1998; Eyer, 1995; Steenland, 1996; Ray, 1998a, b). In one review, these effects were all ranked under the heading of chronic OP-induced neuropsychiatric disorders (COPIND, phenomenon 2; Jamal, 1997). As previously discussed (see Section 72.1.3.9), the various effects will be discussed

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separately for better comprehension and because there is no evidence that they collectively represent a single nosological entity.

72.4.1.1  Neurological Effects on Central Nervous System Clinical reports A case of parkinsonism was described in a subject with reported past and prolonged exposure to OPs. Apparently he also had several episodes of acute poisoning with parathion and malathion said to have required treatment with oral doses of atropine to control symptoms and signs (Davis et al., 1978). The past history was not fully reported and it is doubtful that oral atropine would have been effective because it is known to be poorly and unreliably absorbed. Consequently, this case remains an isolated and anecdotal report. A visual syndrome, known as Saku disease, which is characterized by reduced visual field, myopia, astigmatism, lesions of the optic nerve, and abnormal retinal functions, was associated with the extensive use of OPs during the 1960s in one area of Japan (Saku). These effects, exclusively reported by Japanese investigators, have been summarized by Pleština and Piukovic´-Pleština (1978) and Dementi (1994). However, the symptoms were not consistant among various OPs to which patients were allegedly exposed and often, but not always, associated with AChE inhibition. This inconsistency raises the question whether the effects are compound-specific and related to RBC AChE inhibition. These results have been criticized and the etiological link between OPs and Saku disease remains, for the time being, speculative (Erikson-Lamy and Grant, 1992). Veterans who took part in the Persian Gulf War reported higher rates of many symptoms, including neurological ones, and had a decreased perception of well-being (Ismail et al., 1999; NIH, 1994). Although a single consistent pattern of symptoms and signs is far from being defined (Gray et al., 1996), this mysterious ailment is now known as the Gulf War syndrome or illness, and several hypotheses have been made concerning causes (Lotti, 1999). One theory states that wartime exposure to a combination of OPs and other cholinesterase inhibiting chemicals synergistically produced the syndrome and the neurological signs in particular, (Haley and Kurt, 1997). Among these chemicals, pyridostigmine bromide was the only defined risk factor (Shen, 1998) because it was given to soldiers who served in the Gulf, apparently for several weeks, as prophylaxis for possible nerve gas attacks. The dosing regime was a 30 mg tablet every 8 hours and it aimed to cause reversible inhibition of AChE at nerve endings, thereby preventing irreversible inhibition of the enzyme by OP weapons. None of the soldiers ever experienced acute cholinergic symptoms and signs compatible with OP or pyridostigmine poisoning.

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In a study based on a questionnaire submitted by 249 Gulf War veterans from a single battalion of 606 soldiers, factor analysis of symptoms yielded several syndrome factors (possibly variants of a single syndrome) that suggested various neurological dysfunctions (Haley et al., 1997a). Subjects with the highest factor scores on syndrome 1 (impaired cognition), syndrome 2 (confusionataxia), and syndrome 3 (arthro-myo-neuropathy), for a total of 23 cases, were evaluated for neurological functions and compared with 20 controls from the same battalion, 10 of whom had been deployed in the war region but had no complaints and 10 of whom had not been deployed (Haley et al., 1997b). Brain dysfunction was evidenced by changes in auditory evoked potentials, interocular asymmetry of nystagmic velocity, asymmetry of saccadic velocity, and somatosensory evoked potentials. However, no clinical differences between cases and controls were detected on neurological examination. Exposures to anticholinesterase chemicals of either cases or controls were not reported. Therefore, whether the Gulf War syndrome exists, whether it affects the nervous system, and whether the clinical findings are due to anticholinesterases cannot be ascertained from these studies. Occupational Exposure Studies  Minimal EEG disturbances were reported in a study on 50 workers engaged in the manufacture of a range of unspecified OPs (Metcalf and Holmes, 1969). These changes were not seen in 22 controls and mirrored, to a lesser degree, the more severe disturbances usually seen after acute exposures. Work history and exposure data were insufficient, although it was claimed that the workers were also exposed to chlorinated hydrocarbons. Certain neuropsychological changes were also reported, but it is not clear whether they were associated with such persistent EEG changes. In another study, 32 workers exposed to both OPs and organochlorine (dieldrin) pesticides were subdivided into two equal groups, low and high exposure, based on occupational history. Plasma cholinesterase levels were the same in both groups. EEG and neuropsychological changes were found in the high exposure group (Korsak and Sato, 1977). Quantitative exposure data were not given and EEG changes were different from those reported in the abovementioned study. A selective effect on the left frontal hemisphere, as derived from EEG and neuropsychological results, was detected. This seems inconsistent with a toxic effect. In a seven country biomonitoring and cross-sectional epidemiological survey of low-dose occupational exposure to OPs, changes in EEG were reported only from some countries. Data from one country showed postseason slow wave activity, whereas data from another country reported different changes (Richter, 1993). These results are difficult to assess given the lack of information on exposure and other confounding factors at the time of testing.

Chapter | 72  Clinical Toxicology of Anticholinesterase Agents in Humans

72.4.1.2  Neurological Effects on Peripheral Nervous System Clinical Reports  In a study on volunteers, mevinphos (25 g kg1) was administered daily for 28 days to eight subjects, whereas placebo was given to eight controls (Verberk and Sallé, 1977). RBC AChE was depressed by 19%, but no correlation was found with the detected changes. At the end of exposure, a 7% decrease in slow fiber motor nerve conduction velocity and a 38% increase in Achilles tendon reflex force were found (as percentages of preexposure values). No effects on neuromuscular transmission were detected. The authors concluded that the significance of such effects with regard to health is unclear. Sensory neuropathies on a series of patients with lowlevel exposures to chlorpyrifos have been consistently associated with mild or no cholinergic symptoms (Kaplan et al., 1993) although there is little evidence, if any, for a causal relationship between sensory neuropathy and lowlevel exposures to chlorpyrifos (see Section 72.3.3.1). In a pilot study, 14 Gulf War veterans were examined for peripheral nerve dysfunction and compared with a control group. Although differences were detected in some parameters (cold threshold, sural nerve latencies, and median nerve sensory action potential), the authors’ conclusion was that there may be dysfunction in veterans, but more studies are required (Jamal et al., 1996). Moreover, the hypothesis that anticholinesterases, other than pyridostigmine, represented a risk factor for veterans of the Gulf War remains to be demonstrated. A clinical study was performed on 72 selected subjects with long-term exposure to sheep dip OPs identified in pilot field studies (Pilkington et al., 1999). According to defined criteria and neuropathy scores, 23 workers were ranked as having probable/definite neuropathy, 34 workers had possible neuropathy, and 15 workers had no neuropathy. Clinical evaluation, quantitative sensation testing, nerve conduction studies, and electromyography were performed. Results showed neurological signs in 10% of subjects and some small fiber abnormalities in 65% of electrophysiological tests. None of subjects with symptoms was in the no neuropathy subgroup. This neuropathy was thought to be different from classical OPIDP (see Section 72.2), because sensory fiber almost exclusively and small fiber more than large fiber populations were affected. Clinical and electrophysiological assessments were apparently performed at variable times after cessation of peak exposures (up to 1.5 years). Unlike other toxic neuropathies, this apparently new entity also affected upper limbs at early stages, but unlike OPIDP caused by commercial OP pesticides, there was no preceding cholinergic toxicity. Results from this study are difficult to evaluate because they are not presented analytically. Electrophysiological changes were overwhelmingly more frequent than neurological

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ones, but it is impossible to derive what electrophysiological abnormalities were found in patients with clinically detectable neuropathy. Moreover, the relevance of electrophysiological alterations that are not associated with signs and symptoms is unclear. According to the criteria of sample selection, a causal relationship with exposure cannot be inferred; there is also a lack of relationship with the estimated cumulative dose. Finally, we should ask why, after such a long period since cessation of exposure, a very mild neuropathy did not recover. It seems that perhaps the study better represents a validation of a screening system to detect minor electrophysiological signs to be used in the field than a demonstration of a causal relationship between low-level exposure to OPs during sheep dipping and the development of a new form of toxic peripheral sensory neuropathy. Occupational Exposure Studies  Neuromuscular function was assessed with surface electrodes on the upper limbs of workers exposed to OPs and organochlorine pesticides (Jager et al., 1970). A higher incidence of electromyographic changes (repetitive activity and reduced amplitude) was detected in workers exposed to both chemical classes (n  36) as compared to those exposed to organochlorine only (n  24) and controls (n  28). The biological significance of these small changes is unclear, in part because of the use of surface electrodes. There was insufficient information, only statements, concerning exposures. Changes were thought to be related to synaptic dysfunction because they were similar to changes found in myastenic patients who were overtreated with anticholinesterase drugs. However, when observed in such patients, these changes are associated with substantial inhibition of AChE, whereas changes in workers were not associated with whole blood AChE inhibition. Fifty-three workers exposed to both OPs and organochlorine were examined shortly after the start and toward the end of a spraying season (Drenth et al., 1972). Surface electrode electromyography records of 12 subjects changed from normal to abnormal, whereas those of 13 subjects changed from abnormal to normal. No differences in blood AChE were detected. No evidence of exposure was given. Therefore, the conclusion of the authors that electromyography abnormalities represent only an indication of the need for more protection of workers and no evidence of immediate health problems is not substantiated. Minimal electromyographic changes were detected by surface electrodes in 102 workers exposed to OPs when compared to an unmatched control group of 75 subjects (Roberts, 1976). Fifty-six workers were examined before and after a holiday period. Subjects who displayed these changes somewhat improved after the holidays, whereas some unspecified variability was observed in exposed subjects with normal electromyography. No exposure data were available.

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In a longitudinal study on six workers exposed to OPs over a 7–9 month period, surface electrode electromyography indicated that voltages varied in a manner that reflected a vague assessment of the pattern of exposure (Roberts, 1977). It is difficult to evaluate these results given the lack of exposure data and the methods used. Neuromuscular function was assessed with surface electrodes in a group of 11 spraymen exposed to OPs (including bromophos, diazinon, chlorpyrifos, and malathion) on a recurrent basis (Stålberg et al., 1978). Plasma cholinesterases were significantly reduced after work, whereas RBC AChE was not. A slight reduction in sensory conduction velocity and increased fiber density was detected in some workers, but was unrelated to lowered plasma cholinesterase activity. Although exposure was not assessed in this study, plasma enzyme inhibition indirectly suggests a lack of correlation between degree of exposure and detected electrophysiological changes. Another study was conducted on four groups of subjects in which approximate exposure was assessed: 42 highly exposed to OP pesticides, 14 seasonal workers exposed to OPs and reexamined after exposure, 129 agricultural workers with low exposure to OPs, and 26 agricultural workers not exposed (Jušić et al., 1980). The authors concluded that synapse testing with needle electromyography and clinical examination did not detect latent OP intoxication. A study was conducted on workers exposed to the defoliant DEF, where needle electromyography and biochemical studies (lymphocytic NTE and blood cholinesterases) were performed before and after the spraying season (Lotti et al., 1983). Air and dermal exposure were measured on a typical working day. No electrophysiological changes were detected, although lymphocytic NTE was about 60% inhibited. NTE inhibition after exposure without correlation with electrophysiological changes was explained by the pharmacokinetics of DEF, which requires metabolic activation to be an esterase inhibitor and occurs mainly in the liver; the active metabolite formed is extremely reactive and unless large amounts are formed, it will not reach the nervous system, but will reach the blood, where it inhibits lymphocytic NTE. Twenty-four workers exposed to fenthion were examined with surface electromyography before and after exposure and compared with 19 unexposed controls (Misra et al., 1988). Serum AChE was also measured. Electrophysiological findings after exposure were no different from controls. However, mean values of some electrophysiological parameters were altered in the exposed group when results obtained during exposure were compared with follow-up data collected 3 weeks after withdrawal from exposure. Also, mean cholinesterase values increased after the end of exposure, but remained within the normal range. Although each exposed individual was his or her own control, results are difficult to interpret because the intraindividual variability of measured parameters of controls was not reported.

Hayes’ Handbook of Pesticide Toxicology

Two-hundred twenty-nine workers at a pesticide plant were examined clinically (for neurological impairment) and biochemically (lymphocytic NTE and plasma cholinesterase), and tested for tactile sensitivity and motor performance (Otto et al., 1990). These workers were engaged in the production of a variety of OPs including diazinon, dimethoate, malathion, phentoate, EPN, leptophos, methamidophos, and trichorfon. Results were compared with those obtained from 180 workers from a fertilizer plant and 167 workers from a textile plant. Mean serum cholinesterases and lymphocytic NTE were lower in pesticide workers, although they were within normal ranges. The proportion of workers with abnormal neurological findings (involuntary tremors and vibration sense) varied between plants. Tactile thresholds in the finger of the nondominant hand were higher in workers in the pesticide plant and the authors stated that this symptom was the most sensitive index of pesticide neurotoxicity. Toes were not tested. No changes were detected in the neurobehavioral tests. Assessment of exposure was missing (although some OPs that potentially cause OPIDP were manufactured), the incidence of various diseases, including neurological ones, was particularly high (in both cases and controls), and the fact that upper limb neuropathy is not expected in mild OPIDP creates problems in interpretating results. An epidemiological study on 90 pesticide applicators (Stokes et al., 1995) led to the conclusion that prolonged OP exposure is associated with loss of peripheral nerve function. Exposure was assessed by means of urinary excretion of dimethylthiophosphate, one metabolite of azinphos-methyl. However, workers had been exposed to several other OPs and pesticides. The authors’ conclusion was based on a significant increase in mean vibration threshold sensitivity for applicators’ hands as compared to a matched control group. Feet were not affected. Long-term exposure was determined by questionnaire, but it is unclear whether poisonings had occurred in the past. Subjective symptoms were collected off and on season: headache, weight loss, and nightmares were reported more frequently among pesticide workers, but only headache was statistically increased during the season. Because toxic polyneuropathy does not exclusively affect upper limbs and because workers were exposed to many chemicals, there is no evidence that long-term low-level exposures to OPs cause loss of peripheral nerve functions. A cross-sectional study compared 168 spray operators with long-term exposures to OPs with 84 controls (London et al., 1997). No evidence was found between exposure and loss of vibration sense. However, small differences were found on neurobehavioral test batteries based on information-processing parameters. The authors concluded that there was a small overall evidence of chronic effects of OP exposures, but indicated that exposure misclassification may have contributed to these findings. In another study, the same authors (London et al., 1998) investigated neurological symptoms, vibration

Chapter | 72  Clinical Toxicology of Anticholinesterase Agents in Humans

sense, and tremors in much the same population during the peak spraying season. Eighty-three nonspraying workers were used as the control group. Exposure, as in the previous study, was derived from a job-exposure matrix for pesticides in agriculture. Applicators significantly reported more dizziness, sleepiness, and headache, and had a higher overall neurological symptom score. Vibration sense and tremor outcome were not associated with past long-term OP exposure. A correlation was found between symptoms and either current exposure or episodes of past OP poisoning (see also Section 72.1.3.9). The effects of low-level exposure to foliar OP residues (primarily to azinphos methyl and possibly to phosmet and methyl-parathion) during one season were assessed in a cross-sectional study on 67 workers and 68 matched controls (Engel et al., 1998). Sensory and motor nerve conduction velocities, neuromuscular junction testing, and RBC AChE were measured. No differences were found between exposed and controls.

72.4.1.3  Psychiatric Effects Clinical Reports  Schizophrenic and depressive reactions, with severe impairment of memory and difficulty in concentration were reported in 16 workers after variable exposures to OPs (Gershon and Shaw, 1961). This report was criticized because of serious flaws, including the lack of evidence for exposure, the detailed clinical description of only a few cases, and the inconsistency with larger studies (Barnes, 1961; Bidstrup, 1961). An anecdotal report suggested a causal relationship between psychiatric disturbances and exposure to a variety of pesticides including OPs in two pilots (Dille and Smith, 1964). Another anecdotal report linked the onset of psychosis in a farmer with previous spraying of demeton-S-mefhyl, but no casual relationship was established (Bradwell, 1994). Geographical Studies  A geographical study was carried out to determine whether areas of high OP usage in Australia had a higher proportion of admissions for psychiatric disorders than low-usage areas (Stoller et al., 1965). No evidence was found that schizophrenia, depressive states, psychoneuroses, or personality disorders were more common in high-usage areas than elsewhere. Increased risk of suicide was associated to pesticide exposure (mainly OPs) in an agricultural area (Parrón et al., 1996a). The rate of suicides was compared with rates where exposure to pesticides was low. Most suicide cases involved pesticides, but other factors that influenced suicide attitudes were not analyzed. Occupational Exposure Studies  Workers who had unspecified exposures to OPs were compared with a control group on personality tests, structured interview, and cholinesterase levels (Levin et al., 1976). Commercial sprayers, but not farmers, showed higher levels of anxiety and lower plasma cholinesterase than controls. The authors

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concluded that these findings were tentative until confirmed by additional studies. The effect of exposure stress in the absence of exposure was reported during a manufacturing accident with malathion (Markowitz et al., 1986). The reactions of allegedly exposed workers were compared with a matched group. The exposed group showed high demoralization scores, particularly among those who admitted to little knowledge about toxic chemicals. Twenty-five greenhouse workers were compared to controls and showed a higher incidence of symptoms of depression and tremors (Parrón et al., 1996b). Exposure was not measured and blood cholinesterases were normal. Two groups, one of pesticide formulators (208 individuals) and another of applicators (172 individuals), were compared with matched controls (72 and 151 individuals, respectively; Amr et al., 1997). Exposures to a variety of pesticides, including OPs, carbamates, organochlorine, and pyrethroids, were not quantified. Both exposed groups had a higher incidence of total psychiatric disorders, whereas formulators had a higher incidence of depressive neurosis that was related to the duration of employment. It is difficult to assess the role, if any, of the OPs, given the variety of pesticides to which these workers were exposed. A case control study investigated the link between exposure to pesticides and suicide in Canadian farmers (Pickett et al., 1998). Results excluded exposure to pesticides as an important risk factor for suicide among farmers. However, the chemicals used were not identified and were only divided between herbicides and insecticides. The latter certainly included OPs. Therefore, it cannot be ascertained from this study whether exposures to OPs were involved.

72.4.1.4  Neurobehavioral Effects The neurobehavioral effects of long-term exposures to low levels of OPs have been extensively reviewed over the last few years (D’Mello, 1993; ECETOC, 1998; Eyer, 1995; Jamal, 1995; Mearns et al., 1994; Ray, 1998a, b; Steenland, 1996). Although much information has been published, results are contradictory and whether such exposures are linked with an increased risk of behavioral effects in humans is controversial. A study compared two groups of 53 and 68 asymptomatic workers with varying degrees of unquantified and unspecified exposure to OPs to controls (25 and 22 subjects, respectively) on a complex reaction time test. Results showed there was no indication that exposure at levels insufficient to produce clinical illness had any important effect on mental alertness (Durham et al., 1965). Another study selected 23 workers who regularly used OPs and had used them within 2 weeks of the testing date (Rodnitzky et al., 1975). Recent exposure was confirmed by lower plasma cholinesterase, but RBC AChE was normal. Types of pesticides were not reported. The results of

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tests for memory, signal processing, vigilance, language, and proprioceptive performance were no different from those of a matched control group. Neurobehavioral tests were performed before and after work shifts on 99 pest control workers with low-level, shortterm exposure to diazinon (Maizlish et al., 1987). Exposure was measured by means of the urinary metabolite diethylthiophosphate before and after the end of shifts. No changes in neurobehavioral functions were detected on a battery of seven tests. Similarly, no changes were seen when workers were subdivided according to the degree of exposure. Neuropsychological performance was assessed by test battery in a group of 49 pesticide applicators prior to and 1 month after the end of a 6-month pesticide spraying season. Results were compared with 40 controls (Daniell et al., 1992). The nature and extent of pesticide exposure were assessed and reported in another paper (Karr et al., 1992). The comparison of seasonal RBC AChE changes according to exposure levels showed lower cholinesterase among higher exposure groups compared with lesser exposure group. No evidence of significant decrements in neuropsychological performance was reported. The neurobehavioral status was assessed in workers and kibbutz residents differently exposed to OPs and other pesticides (Richter et al., 1992). Subjects were examined during the spraying season and afterward. Most neurobehavioral scores were poorer during the season. Exposure data were not reported and the authors drew attention to other risk factors such as work load and heat stress. In a cross-sectional study, neuropsychological performance in 146 sheep farmers was compared to 143 quarry workers (Stephens et al., 1995a, b). The selection and the testing procedures for workers who belonged to the experimental group were different from those of controls. Long-term exposure data were assessed by means of a retrospective exposure questionnaire that used the number of sheep, dips, and years of employment as a surrogate. Farmers performed significantly worse than controls in tests to assess sustained attention and speed of information processing (simple reaction time, symbol digit substitution, and syntactic reasoning). A dose–response relationship was found only for one test (syntactic reasoning). Moreover, in another article, no association was found between the experience of acute symptoms and performance on neuropsychological tests (assessed on a subset of workers), and it was concluded that neuropsychological data reflect chronic effects that occur independently of acute effects (Stephens et al., 1996). Given the large differences between OP and quarry workers, it is doubtful that the small changes detected in the former should be attributed to low-level long-term exposures to OPs. Fifty-seven licensed applicators were compared on several neuropsychological tests to a control group of 34 farmers who had no exposure to pesticides (Fiedler et al., 1997). Exposure to OPs was assessed with a questionnnaire on

Hayes’ Handbook of Pesticide Toxicology

work history, but details were not given. None of the applicators had episodes of acute poisoning, and RBC AChE values were normal. Except for slower reaction time, no other difference in neuropsychological performances was detected between exposed and nonexposed subjects. Subclinical morbidity patterns, including symptoms, aiming at digit symbol tests, and measurement of RBC AChE, were investigated in 226 established farm workers and were compared with an equal number of controls and with 92 new farm workers (Gomes et al., 1998). Results indicated a higher incidence of symptoms (irritated conjunctiva, watery eyes, blurred vision, dizziness, headache, and muscular pain and weakness), reduced performance on the aiming and digit symbol tests, and reduced AChE activity in the group of established farmworkers. Although RBC AChE inhibition implies OP exposure, no actual exposure data were reported. Moreover, reduction of AChE was still within the coefficient of variation of the test. Nevertheless, because the above-reported symptoms are consistent with cholinergic overstimulation, it is likely that differences between groups reflect the effects of current exposures to OPs.

72.4.2  Other Effects Several toxic effects have been associated with long-term exposures to OPs. However, most of them are either isolated reports or are based on circumstantial evidence of exposure and probably are simply coincidental. A case report, for instance, suggested congestive cardiomyopathy caused by long-term exposure to OPs without signs of severe acute poisoning. Evidence of exposure was limited and the patient had a previous myocardial infarction, therefore, cardiomyopathy was likely a consequence of myocardial infarction rather than OP exposure (Fazekas and Kiss, 1980). Several cases of influenza-like symptoms associated with OP use in farmers during the sheep dipping season apparently were unrelated to RBC AChE inhibition (Murray et al., 1992). Contact dermatitis and asthma have been linked to exposures to OP pesticides (Bryant, 1985; Deschamps et al., 1994; Xue, 1992). Whether the effects are due to sensitization or irritation, or if they are linked to other ingredients always present in commercial formulations of pesticides is unclear. Immunotoxicity of OPs has been suggested, although human data are mostly based on in vitro studies (Newcombe and Esa, 1992; Rodgers et al., 1992; Sharma and Tomar, 1992). Hypotheses propose that exposures to OPs are linked to cancer development (Newcombe, 1992). One study was performed on workers engaged in the production of several OPs (trichlorfon, chlorfenvinphos, malathion, dichlorvos, fenitrothion, and formothion; Hermanowicz and Kossman, 1984). Exposure involved other chemicals and its assessment was rather approximate. RBC AChE and plasma ChE

Chapter | 72  Clinical Toxicology of Anticholinesterase Agents in Humans

were reported on a group basis and correlated with the assessment of exposure. A marked impairment in neutrophil chemotaxis was found in workers who were likely to be exposed to OPs. The frequency of upper respiratory tract infections was higher in the exposed group compared with controls. No other types of infections showed increased frequency. The authors themselves concluded that a distinction cannot be made between OPs and other chemicals as possible factors for the described effects. Certainly more information is needed to ascertain whether immunotoxicity is an effect of OPs and what pathophysiological significance should be attributed to it. There is no evidence so far that any form of immunomediated clinical effect is linked to OP exposures.

72.4.3  Biomonitoring Occupational Exposures Long-term occupational exposures to OPs are commonly monitored by measuring either urinary excretion of alkylphosphates or blood cholinesterases. The goal is to prevent adverse effects from OPs.

72.4.3.1  Assessment of Urinary Alkylphosphates Although OPs may be excreted unchanged, they are usually hydrolyzed, and the acidic and alcoholic moieties can be found in the urine of exposed subjects. Measurement of metabolites is common practice in workers exposed to OPs, and several alkylphosphates have been identified, including dimethylphosphates, dimethylthiophosphates, dimethyldithiophosphates, and dimethylphosphorothioates derived from dimethylated OPs and the corresponding metabolites derived from diethylated OPs (Coye et al., 1986). Several gas-chromatographic methods to measure urinary dialkylphosphates have been developed (Aprea et al., 1996; Nutley and Cocker, 1993). Measurement of excretion of the alcoholic moiety in exposed workers has been used less frequently. Examples include the measurement of 3, 5, 6-trichloropyridinol after exposure to chlorpyrifos and chlorpyrifos-methyl (Nolan et al., 1984), p-nitrophenol after exposure to parathion and parathion-methyl (Morgan et al., 1977), and mono- and dicarboxylic acid after exposure to malathion (Bradway and Shafik, 1977). Despite the numerous field studies where exposures have been assessed by means of urinary metabolites, not many data are suitable for a toxicological interpretation. The reasons are many. The first problem is related to usual agricultural practices, which lead to concurrent exposures to several OPs. As stated earlier, different OPs, each with its own toxicity, may be metabolized, yielding the same product. It is, therefore, difficult to assess the toxicological

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risk associated with such exposures. For instance, certain exposures to either parathion-methyl or chlorpyrifosmethyl caused comparable excretion of dimethylphosphates and dimethylthiophosphates. However, the risk derived from each OP is quite different, because they display a 3 orders of magnitude difference in acute toxicity (Moretto et al., 1995). Depending on the OP, route of exposure, metabolism, and distribution, peak metabolite excretion might be reached at different times after the end of exposure. For instance, certain compounds such as chlorpyrifos show peak urinary excretion of ethylphosphates several hours after the end of exposure (Fenske and Elkner, 1990; Moretto et al., 1995). On the contrary, peak excretion of ethylphosphates derived from exposures to parathion occurs within a much shorter time (Morgan et al., 1977). Moreover, alkylphosphates may have a different time course of excretion. For instance, diethylphosphate peaks earlier than diethylphosphorothioate after exposure to diazinon (Sewell et al., 1999). Therefore, the timing of urine sampling is crucial to assess the significance of a given concentration; for many OPs, the relevant timing information is missing. Finally, very little is known about the correlation between urinary metabolite excretion and the inhibition of AChE and/or plasma BuChE. In many cases, enzyme inhibition was not found (Griffin et al., 1999; Kraus et al., 1977, 1981; Krieger and Thongsinthusak, 1993; Popendorf et al., 1979); in a few cases, minimal inhibition was found (Jauhiainen et al., 1992; Spear et al., 1977). Only at a time when occupational exposures were probably much more severe than nowadays, was a good correlation found between p-nitrophenol and RBC AChE inhibition in parathion exposed workers (Arterberry et al., 1961). In conclusion, for the time being, it is difficult to give these data a toxicological significance beyond that of being a qualitative exposure index.

72.4.3.2  Monitoring Blood Cholinesterase Blood cholinesterase activities have been used extensively to monitor the effects of occupational exposures to OPs. Guidelines have been developed on methods, interpretation of results, and actions to be taken (EPA, 1992; Pleština, 1984; WHO, 1986). However, these suggestion should be taken as general indications, particularly when interpreting single data, because the following issues must be considered (Lotti, 1995). Relationship between Inhibition of Blood Enzymes and Cholinergic Toxicity The postulate for using blood cholines-terases to biomonitor occupational exposures to OPs is that inhibition of these enzymes reflects either or both the degree of exposure or the corresponding enzyme inhibition in the nervous tissues. Because no physiological functions have been attributed to BuChE (confirmed

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by the fact that homozygote carriers of defective BuChE are healthy subjects; see succeeding text), the inhibition of this enzyme in any tissue most likely has no significance in terms of health. However, its inhibition in plasma means that exposure has occurred. This statement may not be true in the case of diseases (unlikely in occupational exposures) that cause depression of plasma BuChE, such as parenchimal liver diseases, acute infections, some anemias, and malnutrition. In this respect, an interesting observation is that patients with liver diseases not only have low plasma cholinesterases, but also may show a further reduction as a result of a level of exposure to OPs that causes no change of enzyme activity in normal persons or in persons affected by other diseases (Hayes, 1982). A possible explanation is the reduced ability of patients affected by liver diseases to detoxify certain OPs. Obviously, if BuChE inhibition is associated with inhibition of RBC AChE, then different conclusions should be drawn. How accurately inhibition of RBC AChE reflects that in the synapses is unknown and extrapolation is difficult, given the different access various OPs have to the blood and the nervous system. Animal data suggest that sometimes inhibition is similar, more often, inhibition of blood enzyme is higher (Su et al., 1971) due to the particular protection of the nervous system that is offered by the blood–brain barrier. After recovery from exposure, extrapolation is even more difficult, given the different rates of recovery of AChE in the RBCs and nervous tissues, respectively. Nevertheless, based on clinical and toxicological data, a rough estimate of the levels of RBC AChE inhibition that require action are reported in the Table 72.10. It is also clear that because the access of xenobiotics to blood is always easier than to brain and because no evidence exists that OPs accumulate in the nervous system, the inhibition of RBC AChE usually overestimates the level in the brain. Spontaneous Reactivation and Reappearance of Blood Enzymes When measuring blood cholinesterase for biomonitoring purposes, the rates of reappearance of activities after inhibition should be taken into account. Such rates depend on spontaneous reactivation (in the case of carbamates and dimethylphosphates) and resynthesis of new enzymes. As previously discussed (Section 72.1.2), the rate constants of spontaneous reactivation and of aging vary according to the phos-phorylating agent. Because rate constants also depend on the enzyme that is phosphorylated, they will be different when measured in RBC AChE or plasma BuChE (Aldridge and Reiner, 1972). However, the few studies in humans do not necessarily confirm these theoretical considerations. Data from school children treated orally with trichlorfon against schistosomiasis showed that plasma BuChE as well as RBC AChE recovered much slower than was predicted from in vitro spontanous reactivation studies (Reiner and Pleština, 1979).

Table 72.10  Relationship between RBC AChE Inhibition and Preventive Actions When Monitoring Occupational Exposures to OPsa RBC AChE (% inhibition from preexposure values)

Significance

Preventive action

20–29

Evidence of exposure

Improve hygenic conditions

30–50

Hazard

As above plus removal of subject from exposure

50

Poisoning

Admit subject to the hospital

a

Data from WHO (1986) and Lotti (1995).

The synthesis of AChE occurs in the bone marrow, and its presence in the blood depends on the normal turnover of RBCs (i.e., 120 days). The synthesis of BuChE occurs in the liver and its turnover in plasma corresponds to about 20 days. Resynthesis of both enzymes after irreversible inhibition by OPs in the nervous systems of animals seems to occur at similar rates, corresponding to a half-life of 5–7 days. However, resynthesis is reflected in the blood quite differently because the localizations of AChE and BuChE differ. Thus, the reappearance of RBC AChE has been shown to occur at a rate of about 1% per day, whereas the rate of plasma BuChE is about 5% per day (Hayes, 1982). Sensitivity of Blood Enzymes to Inhibitors  As stated previously, RBC AChE and plasma BuChE are different enzymes, and, therefore, they display different substrate specificity. Whereas interactions of OPs with esterases are analogous to interactions of substrates with esterases, it is clear that blood cholinesterases are differently inhibited by a given OP. As shown in Table 72.11, plasma BuChE is generally more sensitive to inhibition than RBC AChE by most OPs used as insecticides. In cases of mild exposures, plasma BuChE may be the only inhibited enzyme. This observation should be interpreted as a sign of exposure, but not of poisoning. In cases of severe exposures, profound inhibition of plasma BuChE is always associated with similar inhibition of RBC AChE as it occurs in poisoned patients. The situation may be more complex when substantial repeated exposures are involved. In such a case, even if the plasma enzyme is more sensitive, a buildup of RBC AChE inhibition can occur, thus equalizing the inhibition of both enzymes at a certain time because of the different rates of reappearance of the two enzymes. Therefore, equal inhibition of the enzymes may represent the consequence of either a single substantial exposure or less severe but repeated exposures. Finally, RBC AChE may be more inhibited, even if the plasma enzyme is more sensitive when subjects are

Chapter | 72  Clinical Toxicology of Anticholinesterase Agents in Humans

Table 72.11  Sensitivity of Plasma BuChE and RBC AChE in Humans to Various Insecticidesa RBC AChE most inhibited

Plasma cholinesterase most inhibited

Dimefox (Edson, 1964)

Chlorfenvinphos (FAO/WHO, 1995)

Mevinphos (Rider et al., 1972, 1975)

Chlopyrifos (Eliason et al., 1969)

Methyl-parathion (Rider et al., 1970)

Demeton (Moeller and Rider, 1965)

Parathion (Hayes, 1982)

Diazinon (FAO/WHO, 1967) Dichlorvos (Rasmussen et al., 1963) Fenitrothion (Vandekar, 1965) Malathion (Elliot and Barnes, 1963) Monocrotophos (FAO/WHO, 1996) Trichlorfon (Abdel-Al et al., 1970)

a

Circumstances of exposure vary.

recovering from substantial exposures, given the prolonged life of the RBCs that carry inhibited AChE. Inter- and Intraindividual Variability of Blood Enzymes  Intraindividual variability of both plasma BuChE and RBC AChE is high. Samples taken at intervals ranging from a few days to several years indicate that the coefficients of variation of both enzymes in unexposed subjects vary from 7 to 11%. In some cases, an intraindividual variability of plasma enzyme up to 100% was detected in the course of several months (Hayes, 1982). Interindividual variability of these enzymes is even greater. The coefficients of variation of RBC AChE in unexposed subjects vary from 10 to 40%, whereas the corresponding values for plasma enzyme vary from 12 to 46%. Minor differences exist according to gender, age, and race (Hayes, 1982). A few people who have normal levels of RBC AChE are genetically deprived of plasma BuChE (Østergaard et al., 1992). It was observed that some of these people who were treated with succinylcholine during surgery exhibited an abnormal prolonged period of muscular paralysis after usual dosages of the drug. These patients were found to have a plasma BuChE much lower than normal (Bourne et al., 1952). Moreover, the enzyme is also qualitatively different from the norm, as for instance, in sensitivity to inhibitors. Because of this, a test was developed based upon lesser inhibition of the enzyme by dibucaine (Kalow and Staron, 1957). The dibucaine number (degree of plasma cholinesterase inhibition by dibucaine) discriminates three phenotypes: normal, intermediate, and atypical. The approximate frequency of these phenotypes has been estimated at 96, 3.9, and 0.05%, respectively (Harris and

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Whittaker, 1962). Other tests are available to discriminate abnormal BuChE, based upon fluoride number (Harris and Whittaker, 1961), chloride number (Whittaker, 1968), scoline number (King and Griffin, 1973), and urea number (Hanel and Viby-Mogensen, 1977; see Section 1.3.6.4).

72.4.3.3  Detection of Hypersusceptible Subjects Whereas OPs are inhibitors of plasma BuChE and are largely hydrolyzed by A-esterases, such as paraoxonase (PON1) and other carboxylesterases (aliesterases), inherited or acquired deficits of scavenger (plasma BuChE) or detoxifying (esterases) abilities have been suggested as potential factors for increased susceptibility to OPs (Loewenstein-Lichtenstein et al., 1996; Saxena et al., 1997). Although there is some evidence in experimental animals for hypersusceptibility based upon reduced ability to detoxify OPs, only one example is known in humans. An unexpected outbreak of malathion poisoning arose in workers occupationally exposed to commercial brands of malathion that contained high amounts of impurities. Among these, isomalathion was the most relevant because it inhibited the carboxylesterases that inactivativate malathion by hydrolyzing its carboxyl–ester linkages (Baker et al., 1978). Recurrent suggestions have been made that genetically determined low levels of plasma BuChE increased the susceptibility to OP toxicity because of a reduced scavenger capability. However, a relationship between abnormal plasma BuChE and hypersusceptibility to OP poisoning has never been reported. Hypersusceptibility to succinylcholine would never have been discovered if some unusual people had not undergone succinylcholine treatment, an event that is probably no more common than heavy exposure to OPs. Moreover, when the biochemical characteristics of the normal and of the genetically determined defective enzymes were compared stoichiometrically with the plasma concentrations of inhibitors, no scavenger functions to plasma BuChE could be detected (Lotti and Moretto, 1995). Based on the polymorphism of PON1 in human populations and the known role of this enzyme in the detoxification of some OPs (Davies et al., 1996; Mueller et al., 1983), it has been inferred that the expression of this enzyme is involved in determining hypersusceptibility to OPs (Mackness et al., 1998). However, so far, proof for this has been obtained only in animals (Costa et al., 1999).

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Sedgwick, E. M., and Senanayake, N. (1997). Pathophysiology of the intermediate syndrome of organophosphorus poisoning. J. Neurol. Neurosurg. Psychiatry 62, 201–202. Selden, B. S. and Curry, S. C. (1987). Prolonged succinylcholine-induced paralysis in organophosphate insecticide poisoning. Ann. Emerg. Med. 16, 215–217. Sellström, Å. (1992). Anticonvulsants in anticholinesterase poisoning. In “Clinical & Experimental Toxicology of Organophosphates and Carbamates” (B. Ballantyne and T. C. Marrs, eds.), pp. 578–586. Butterworth–Heinemann, Oxford. Senanayake, N. (1981). Tri-cresyl phosphate neuropathy in Sri Lanka: A clinical and neurophysiological study with a three year follow up. J. Neurol. Neurosurg. Psychiatry 44, 775–780. Senanayake, N. and Jeyaratnam, J. (1981). Toxic polyneuropathy due to gingili oil contaminated with tri-cresyl phosphate affecting adolescent girls in Sri Lanka. Lancet i, 88–89. Senanayake, N., and Johnson, M. K. (1982). Acute polyneuropathy after poisoning by a new organophosphate insecticide. N. Engl. J. Med. 306, 155–157. Senanayake, N., and Karalliedde, L. (1987). Neurotoxic effects of organophosphorus insecticides. An intermediate syndrome. N. Engl. J. Med. 316, 761–763. Senanayake, N., and Sanmuganathan, P. S. (1995). Extrapyramidal manifestations complicating organophosphorus insecticide poisoning. Human Exp. Toxicol. 14, 600–604. Sewell, C., Pilkington, A., Buchanan, D., Tannahill, S. N., Kidd, M., Cherrie, B., and Robertson, A.(1999). “Epidemiological Study of the Relationships between Exposure to Organophosphate Pesticides and Indices of Chronic Peripheral Neuropathy, and Neuropsychological Abnormalities in Sheep Farmers and Dippers. Phase 1. Development and Validation of an Organophosphate Uptake Model for Sheep Dippers.” Technical Memorandum Series. Institute of Occupational Medicine, Edinburgh. Sharma, R. P., and Tomar, R. S. (1992). Immunotoxicity of anticholinesterase agents. In “Clinical & Experimental Toxicology of Organophosphates and Carbamates” (B. Ballantyne and T. C. Marrs, eds.), pp. 203–210. Butterworth–Heinemann, Oxford. Shen, Z.-X. (1998). Pyridostigmine bromide and Gulf War syndrome. Med. Hypothesis 51, 235–237. Shiraishi, S., Inoue, N., Murai, Y., Onishi, A., and Noda, S. (1983). Dipterex (trichlorfon) poisoning. Clinical and pathological studies in human and monkeys. J. UOEH, 5(Suppl.), 125–132. Sidell, F. R. and Groff, W. A. (1971). Intramuscular and intravenous administration of small doses of 2-pyridinium aldoxime methochloride to man. J. Pharm. Sci. 60, 1224–1228. Sidell, F. R., Groff, W. A., and Kaminskis, A. (1972). Toxogonin and pralidoxime: Kinetic comparison after intravenous administration to man. J. Pharmacol. Sci. 61, 1765–1769. Singh, G., Avasthi, G., Khurana, D., Whig, J., and Mahajan, R. (1998a). Neurophysiological monitoring of pharmacological manipulation in acute organophosphate (OP) poisoning. The effects of pralidoxime, magnesium sulphate and pancuronium. Electroenceph. Clin. Neurophysiol. 107, 140–148. Singh, G., Mahajan, R., and Whig, J. (1998b). The importance of electrodiagnostic studies in acute organophosphate poisoning. J. Neurol. Sci. 157, 191–200. Singh, G., Sidhu, U. P. S., Mahajan, R., Avasthi, G., and Whig, J. (2000). Phrenic nerve conduction studies in acute organophosphate poisoning. Muscle Nerve 23, 627–632. Singh, S., Batra, Y. K., Singh, S. M., Wig, N., and Sharma, B. K. (1995). Is atropine alone sufficient in acute severe organophosphorus poisoning?:

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Sussman, J. L., Harel, M., Frolow, R., Oefner, C., Goldman, A., Toker, L., and Silman, I. (1991). Atomic structure of acetylcholinesterase from Torpedo californica: A prototypic acetylcholine-binding protein. Science 253, 872–879. Suzuki, T., Morita, H., and Ono, K. (1995). Sarin poisoning in Tokyo subway. Lancet 345, 980. Tabershaw, I. R. and Cooper, W. C. (1966). Sequelae of acute organic phosphate poisoning. J. Occup. Med. 8, 5–20. Tafuri, J. and Roberts, J. (1987). Organophosphate poisoning. Ann. Emerg. Med. 16, 193–202. Tanaka, D. Jr. and Bursian, S. J. (1989). Degeneration patterns in the chicken central nervous system induced by ingestion of the organophosphorus delayed neurotoxin tri-ortho-tolyl phosphate. A silver impregnation study. Brain Res. 484, 240–256. Taylor, P. (1996a). Anticholinesterase agents. In “Goodman and Gilman’s the Pharmacological Basis of Therapeutics,” (J. G. Hardman and L. E. Limbird, eds.) 9th ed., pp. 161–176. McGraw–Hill, New York. Taylor, P. (1996b). Agents acting at the neuromuscular junction and autonomic ganglia. In “Goodman and Gilman’s the Pharmacological Basis of Therapeutics” (J. G. Hardman and L. E. Limbird, eds.), 9th ed., pp. 177–197. McGraw–Hill, New York. Thompson, J. W., and Stocks, R. M. (1997). Brief bilateral vocal cord paralysis after insecticide poisoning. A new variant of toxcity syndrome. Arch. Otolaryngol. Head Neck Surg. 123, 93–96. Tomlin, C. D. S.(1997). “The Pesticide Manual,” 11th ed.. British Crop Protection Council, Surrey, UK. Tosi, L., Righetti, C., Adami, L., and Zanette, G. (1994). October 1942: A strange epidemic paralysis in Saval, Verona, Italy. Revision and diagnosis 50 years later of tri-ortho-cresyl phosphate poisoning. J. Neurol. Neurosurg. Psychiatry 57, 810–813. Tracey, J. A., and Gallagher, H. (1990). Use of glycopyrrolate and atropine in acute organophosphorus poisoning. Human Exp. Toxicol. 9, 99–100. Tsao, T. C-Y., Juang, Y-C., Lan, R.-S., Shieh, W.-B., and Lee, C.-H. (1990). Respiratory failure of acute organophosphate carbamate poisoning. Chest 98, 631–636. Tsatsakis, A. M., Aguridakis, P., Michalodimitrakis, M. N., Tsakalov, A. K., Alegakis, A. K., Koumantakis, E., and Troulakis, G. (1996). Experiences with acute organophosphate poisonings in Crete. Vet. Human Toxciol. 38, 101–107. Tush, G. M., and Amstead, M. I. (1997). Pralidoxime continuous infusion in the treatment of organophosphate poisoning. Ann. Pharmacother. 31, 441–444. Vale, J. A., and Scott, G. W. (1974). Organophosphorus poisoning. Guy’s Hosp. Rep. 123, 13–25. Valero, A., and Golan, D. (1967). Accidental organic phosphorus poisoning: The use of propranolol to counteract vagolytic cardiac effects of atropine. Isr. J. Med. Sci. 3, 582–584. Vandekar, M. (1965). “Observations of the Toxicity of Two Organophosphorus and One Carbamate Insecticide in a Village Trial Performed by WHO Insecticide Testing Unit in Lagos During 1964”, WHO Work. Doc. 65/Tox/2.64. U.S. Govt. Printing Office, Washington, DC. Van Meter, W. G., Karczmar, A. G., and Fiscus, R. R. (1978). CNS effects of anticholinesterases in the presence of inhibited cholinesterases. Arch. Int. Pharmacodyn. 231, 249–260. Vasilescu, C. (1982). Neuropathy after organophosphorus compounds poisoning. J. Neurol. Neurosurg. Psychiatry 45, 942. Vasilescu, C., and Florescu, A. (1980). Clinical and electrophysiological study of neuropathy after organophosphorus compounds poisoning. Arch. Toxicol. 43, 305–315.

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Vasilescu, C., Alexianu, M., and Dan, A. (1984). Delayed neuropathy after organophosphorus insecticide (dipterex) poisoning: A clinical, electrophysiological and nerve biopsy study. J. Neurol. Neurosurg. Psychiatry 47, 543–548. Verberk, M. M., and Sallé, H. J. A. (1977). Effects of nervous function in volunteers ingesting mevinphos for one month. Toxicol. Appl. Pharmacol. 42, 351–358. Verpooten, G. A., and De Broe, M. E. (1984). Combined hemoperfusionhemodialysis in severe poisoning: Kinetics of drug extraction. Resuscitation 11, 275–289. Wadia, R. S., Sadagopan, C., Amin, R. B., and Sardesai, H. V. (1974). Neurological manifestations of organophosphorus insecticide poisoning. J. Neurol. Neurosurg. Psychiatry 37, 841–847. Wadia, R. S., Shinde, S. N., and Vaidya, S. (1985). Delayed neurotoxicity after an episode of poisoning with dichlorvos. Neurol. India 33, 247–253. Wadia, R. S., Chitra, S., Amin, R. B., Kiwalkar, R. S., and Sardesai, H. V. (1987). Electrophysiological studies in acute organophosphate poisoning. J. Neurol. Neurosurg. Psychiatry 50, 1442–1448. Wang, A.-G., Liu, R.-S., Liu, J.-H., Teng, M. M.-H., and Yen, M. Y. (1999). Positron emission tomography scan in cortical visual loss in patients with organophosphate intoxication. Ophthalmology 106, 1287–1291. Wecker, L., Mrak, R. E., and Dettbarn, W. D. (1985). Evidence of necrosis in human intercostal muscle following inhalation of an organophosphate insecticide. J. Environ. Pathol. Toxicol. Oncol. 6, 171–175. Weeks, D. B., and Ford, D. (1989). Prolonged suxamethonium-induced neuromuscular block associated with organophophate poisoning. Br. J. Anaesth. 62, 327. Weir, S., Minton, N., and Murray, V. (1992). Organophosphate poisoning in the U.K.: The National Poisons Information Service experience during 1984–1987. In “Clinical & Experimental Toxicology of Organophosphates and Carbamates” (B. Ballantyne and T. C. Marrs, eds.), p. 463–170. Butterworth–Heinemann, Oxford. Weizman, Z., and Sofer, S. (1992). Acute pancreatitis in children with anticholinesterase insecticide intoxication. Pediatrics 90, 204–206. Whittaker, M. (1968). The pseudocholinesterase variants. Differentiation by means of sodium chloride. Ada Genet. 18, 562–566. WHO (1986). “Organophosphorus Insecticides: A General Introduction.” Environmental Health Criteria 63, World Health Organization, Geneva. WHO (1998). “The WHO Recommended Classification of Pesticides by Hazard and Guidelines to Classification 1998–1999.” WHO/ PCS/98.21, World Health Organization, Geneva. Whorton, M. D., and Obrinsky, D. L. (1983). Persistence of symptoms after mild to moderate acute organophosphate poisoning among 19 farm field workers. J. Toxicol. Environ. Health 11, 347–354. Willems, J. L. (1981). Poisoning by organophosphate insecticide: Analysis of 53 human cases with regard to management and drug treatment. Acta Med. Milit. (Belg) 134, 7–14. Willems, J. L., and Belpaire, F. M. (1992). Anticholinesterase poisoning: An overview of pharmacotherapy. In “Clinical & Experimental Toxicology of Organophosphates and Carbamates” (B. Ballantyne and T. C. Marrs, eds.), pp. 536–544. Butterworth–Heinemann, Oxford. Willems, J. L., Langenberg, J. P., Verstraete, A. G., De Loose, M., Vanhaesebroeck, B., Goethals, G., Belpaire, P. M., Buylaert, W. A., Vogelaers, D., and Colardyn, F. (1992). Plasma concentrations of pralidoxime methylsulphate in organophosphorus poisoned patients. Arch. Toxicol. 66, 260–266.

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Willems, J. L., De Bisschop, H. C., Verstraete, A. G., Declerck, C., Christiaens, Y., Vanscheeuwyck, P., Buylaert, W. A., Vogelaers, D., and Colardyn, F. (1993). Cholinesterase reactivation in organophosphorus poisoned patients depends on the plasma concentrations of the oxime pralidoxime methylsulphate and of the organophosphate. Arch. Toxicol. 67, 79–84. Wilson, B. W., Padilla, S., Henderson, J. D., Brimijoin, S., Dass, P. D., Elliot, G., Jaeger, B., Lanz, D., Pearson, R., and Spies, R. (1996). Pactors in standardizing automated cholinesterase assays. J. Toxicol. Environ. Health 48, 187–195. Worek, F., Kirchner, T., Bäcker, M., and Szinicz, L. (1996). Reactivation by various oximes of human erythrocyte acetylcholinesterase inhibited by different organophosphorus compounds. Arch. Toxicol. 70, 497–503. Worek, F., Eyer, P., and Szinicz, L. (1998a). Inhibition, reactivation and aging kinetics of cyclohexylmethylphosphonofluoridate-inhibited human cholinesterases. Arch. Toxicol. 72, 580–587. Worek, F., Widmann, R., Knopff, O., and Szinicz, L. (1998b). Reactivating potency of obidoxime, pralidoxime, HI 6 and HLö 7 in human erythrocyte acetylcholinesterase inhibited by highly toxic organophosphorus compounds. Arch. Toxicol., 72, 237–243. Worek, F., Diepold, C., and Eyer, P. (1999a). Dimethylphosphoryl-inhibited cholinesterases: Inhibition, reactivation, and aging kinetics. Arch. Toxicol. 73, 7–14. Worek, F., Mast, U., Kiderlen, D., Diepold, C., and Eyer, P. (1999b). Improved determination of acetylcholinesterase activity in human whole blood. Clin. Chim. Acta 288, 73–90. Xintaras, C., and Burg, J. R. (1980). Screening and prevention of human neurotoxic outbreaks: Issues and problems. In “Experimental and

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Chapter 73

Application of Physiologically Based Pharmacokinetic/Pharmacodynamic Modeling in Cumulative Risk Assessment for N-Methyl Carbamate Insecticides Xiaofei Zhang1, James B. Knaak2, Rogelio Tornero-Velez3, Jerry N. Blancato4 and Curtis C. Dary5 1

General Dynamics Information Technology, Henderson, Nevada The State University of New York at Buffalo, Buffalo, New York 3,4 U.S. Environmental Protection Agency, Research Triangle Park, North Carolina 5 U.S. Environmental Protection Agency, Las Vegas, Nevada 2

73.1  Need for cumulative risk assessments for n-methyl carbamate insecticides Human exposure to xenobiotics may occur through mul­ tiple pathways and routes of entry punctuated by exposure intervals throughout a work or leisure day. Exposure to a single environmental chemical along multiple pathways and routes (aggregate exposure) may have an influence on an organism if the exposure dose is absorbed and dis­ tributed to target tissues. Exposure to multiple chemicals belonging to a class of similar compounds having a com­ mon mechanism of action may have a cumulative effect on target tissues (cumulative toxicity). Under this situa­ tion, the evaluation of the toxic effect from one chemical is obviously not enough. Instead, the net effect from all chem­ icals should be considered. Whenever such scenarios are encountered, cumulative risk assessment (CRA) is needed in order to evaluate the net cumulative toxicity caused by the aggregate exposure from all routes of entry for a sin­ gle chemical or a group of chemicals that have a common mechanism of toxicity (U.S. EPA, 2002a). The application of pesticides for the purpose of insect pest control creates such possible scenarios, not only in occu­ pational settings but also in the general population. Organo­ phosphorus compounds, N-methyl carbamates (NMCs), and pyrethroids are three popular classes of insecticides widely used in the United States and worldwide. For the general population, these pesticide residues may enter Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

exposure pathways in food, drinking water, breathable air, and on residential surfaces where exposure may occur by ingestion, inhalation, and dermal contact. Recognizing the risk imposed by exposure to multiple chemicals, the Food Quality Protection Act (FQPA) of 1996 in the United States and Regulation (EC) No. 396/2005 in the European Union both mandate CRAs on human health resulting from exposure to multiple chemicals that exert their toxicity by a common mechanism of action. The FQPA requires that the U.S. Environmental Protection Agency (EPA) consider the cumulative effects to human health that can result from pesticides and other substances with the same mechanism of toxicity. To know the background and history of regula­ tions regarding CRA, readers are directed to the U.S. EPA (2002b) for the available papers. As insecticides, N-methyl carbamates (NMCs) share a common chemical structure with the general ­ formula ROC(O)NHCH3 for N-methyl carbamates and ROC(O)N (CH3)2 for dimethyl carbamates. The detailed chemical structures of each member in this class can be found in Table 73.1. NMCs have a common mechanism of action toward insect pests and unintended toxicity to nontarget organisms including humans, that is, acetylcholinesterase (AChE) inhibition by carbamylating the serine hydroxyl group in the active site of the enzyme in the nervous sys­ tem, leading to the persistent action of the neurotransmit­ ter, acetylcholine, on cholinergic postsynaptic receptors (Baron, 1991; Ecobichon, 1991; Knaak et al., 2008; Kuhr and Dorough, 1976; Matsumura, 1985; O’Brien, 1967). 1591

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Table 73.1  General Information on the Carbamate Insecticidesa Common name

N-methyl carbamate class

Molecular weight

CAS no.

Aldicarb

oxime

190.26

116-06-3

Chemical structure

O

S N

N H

O

Carbaryl

aryl

201.20

63-25-2

HN

O O

Carbofuran

aryl

221.25

1563-66-2 O O

O NH Carbofuran

Formentanate

aryl

221.26

22259-30-9

N N

O

O HN

Methiocarb

aryl

225.31

2032-65-7

O

HN O

S

Methomyl

oxime

162.21

16752-77-5

O O S

Oxamyl

oxime

219.26

23135-22-0

O O

Pirimicarb

aryl

238.29

N H

N

S

N

O

N

N H

23103-98-2

N

O

N

Propoxur

aryl

209.24

N

O

114-26-1 O

HN O

O

Thiodicarb

oxime

354.47

59669-26-0 N O

S

O N

S

N

N

O O

a

Revised from Knaak et al. (2008).

S

N

Chapter | 73  Application of Physiologically Based Pharmacokinetic/Pharmacodynamic Modeling

Therefore, these pesticides are recognized as a common mechanism group (CMG) (U.S. EPA, 2007). Unlike the organophosphorus insecticides, inhibition of cholinesterase by NMCs is reversible and the onset and recovery of inhib­ ition are rapid, with the maximum inhibition occurring between 15 and 45 min and recovery starting from minutes to hours (U.S. EPA, 2007).

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the reference value for each individual chemical. This method uses the basic principle for risk assessment of ­calculating the hazard index, which is the inverse of the margin of exposure. Readers are referred to the detailed description for various index methods presented in Boobis et al. (2008). It should be noted that, even though the algorithms are different, these methods are interchangeable when the evaluation of all chem­ icals is based on the same toxicological endpoint and study design; thus the outcome of the assessment is the same regard­ less of the method used (Boobis et al., 2008). The RPF and PBPK/PD modeling methods will be described further in the following sections with particular emphasis on the PBPK/PD modeling approach. The use of PBPK modeling for the CRA of pesticides has been discussed or performed by Lowit et al. (2004), Conolly et al. (2005), and Zhang et al. (2008). The EPA has issued a report providing guidance for using the PBPK method in risk assessment (U.S. EPA, 2006a).

73.2  Methodologies for performing cumulative risk assessments for n-methyl carbamates A CRA begins with the identification of a CMG of chem­ i­cals, which exert toxic effects by a common mechanism of action (U.S. EPA, 1999). After the identification of a CMG, individual chemicals are selected based on perceived risk and exposure potential. This subgroup of chemicals within the CMG is sorted into cumulative assessment groups. Many approaches have been investigated and used in CRA (Boobis et al., 2008). Basically, there are four methodolo­ gies that include (1) a toxicological index method, (2) a margin of exposure method, (3) a relative potency factor (RPF) method, and (4) physiologically based pharmacoki­ netic/pharmacodynamic (PBPK/PD) modeling. Algorithms used in the first three methods are summarized in Table 73.2. The index method accounts for cumulative risk by sum­ ming all risk indexes calculated as the ratio of exposure level to

73.2.1  Relative Potency Factor Approach Using an Index Chemical The EPA developed the RPF approach, which uses an index chemical to carry out the CRA for organophospho­ rus insecticides (U.S. EPA, 2006b) and NMCs (U.S. EPA, 2007) and has released guidance for performing CRAs for aggregate exposures involving multiple routes for a single chemical (aggregate risk) (U.S. EPA, 2001) and for a CMG that incurs cumulative risk (U.S. EPA, 2002a).

Table 73.2  Non-PBPK-Modeling Methods for the CRA for the Compounds in the Same Cumulative Assessment Groupa Methodology

Criteria used for CRA

Algorithm

Index method

Hazard index (HI)

Index method

Cumulative risk index (CRI)

CRI 

1 Exp1/ RV 1  Exp2 / RV 2  Exp3 / RV 3  …

CRI  1

Index method

Reference point index (RPI)

RPI 

Exp1 Exp2 Exp3   … RP1 RP 2 RP 3

RPI  1

Margin of exposure

Combined margin of exposure (MOET)

MOET 

Potency factor

Relative potency factor (RPF)b

MOE 

HI 

Acceptable level

Exp1 Exp2 Exp3   … RV 1 RV 2 RV 3

1 (1/ MOE1)  (1/ MOE 2)  (1/ MOE 3)  … PoDindex

HI  1

MOET  100

MOE  10

n

∑ Residue i  PFi  RPFi i 1

a

Summarized from Boobis et al. (2008). RPF method converts other chemicals to the equivalent doses of the selected index compound to derive the total equivalent exposure. MOE, margin of exposure; RV, reference value; PF, processing factor; PoD, point of departure.

b

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Briefly, an index chemical is selected based on the com­ pleteness and representativeness of available data and is used as the point of reference for comparing the absorbed dose for the rest of the chemicals in the cumulative assess­ ment group. The doses for the other CMG chemicals are converted into the equivalent dose of the index chemical so that the aggregate exposure can then be lumped together as the equivalent dose of the index chemical. The cumulative risk is then evaluated by comparing the level of exposure against the point of departure (PoD) on the dose–response curve of the index chemical. The PoD is defined as a dose–response point that marks the begin­ ning of a low-dose extrapolation from laboratory animals to humans. Under most situations, the PoD is based on an exter­ nal exposure or administered dose that leads to the observed responses. The selected PoD is used to depart from the observed range of empirical response data for extrapolating risk in laboratory animals to the exposure anticipated in the human population (U.S. EPA, 2002a). The PoD that is usually used is either a no observed adverse effect level (NOAEL), lowest observed adverse effect level (LOAEL), or benchmark dose (BMD) of the index chemical. The EPA prefers the use of BMD rather than NOAEL or LOAEL (U.S. EPA, 2007). For the NMCs, the endpoint of relative potency is brain AChE inhibition measured at peak inhibition following oral gavage exposures in Long Evans rats (Padilla et al., 2006) and studies submitted by the registrants. The central estimate of 10% brain AChE inhibition (BMD10) is used as the response level to develop RPFs. In the family of NMCs, oxamyl is used as the index chemical because of the availability of high-quality dose–response data for vari­ ous routes. Brain AChE activity was selected as the end­ point to calculate RPFs from the PoDs (U.S. EPA, 2007). Based on the available brain AChE activity data in the rat, an empirical dose–time–response exponential model was developed for each NMC, and the central estimate of the BMD10 is used to determine the relative potency. The lower confidence limit of the BMD10 (i.e., BMDL10) for the index chemical, oxamyl, is used as the PoD. The math­ ematical equations describing the dose–time–response exponential model can be found in U.S. EPA (2007). An RPF is defined as the ratio of the BMD10 of oxamyl divided by the BMD10 of each NMC. With this algorithm, the RPFs for all 10 NMCs range from 0.02 (pirimicarb) to 4 (aldicarb) with the RPF of the index chemical as 1. The exposure doses are then converted to an equivalent dose of oxamyl by multiplying the estimated dose of each NMC with its RPF to calculate the index equivalent residue [used in Eq. (1)]. The aggregate exposure is then summed together to obtain the total exposure of oxamyl as indicated in the denominator in Eq. (1). After further uncertainty factors are applied, the PoD is adjusted to extrapolate the exposure dose for humans. The targeted acceptable margin of exposure [as calculated in Eq. (1)] for the NMC CRA is larger than 10.

MOE 

PoD index

(1)

n

∑ residuei  PFi  RPFi i1



where MOE is the margin of exposure and PF refers to the chemical-specific processing factor. The RPF method is based on several assumptions. First, the dose of each chemical is assumed to be additive; that is, there is no synergism or competitive interaction from all AChE inhibitors in the process of AChE inhibition. This assumption is based on a multi-NMC mixture study (Padilla et al., 2007) in which interactions among NMCs were not observed. Second, the RPF method largely depends on the availability and quality of the toxicological database on the index chemical. Consequently, any uncertainty or incom­ pleteness in this database would be propagated in the down­ stream algorithm. Meanwhile, the assumption of the ratio of toxic potencies between the index chemical and the specific NMC as being constant across all dose ranges needs further experimental testing. Moreover, the aggregated exposure (summed dose) is assumed to occur as a single time event rather than a series of events, and not as discrete or punc­ tuated events, which can actually happen in reality. Lastly, subject differences (age, gender, and metabolic polymor­ phism) are not considered. Therefore, the RPF method can not reflect the variation in the targeted human population. With the advances in PBPK modeling described in the next section, however, these assumptions can be tested.

73.2.2  Advances in the PBPK/PD Modeling Approach Theoretically, PBPK models can be regarded as the “electronic copy” of the laboratory animal or human test sys­ tem. “Exposure” can be simulated in silico and tissue dosim­ etry can be estimated or predicted prior to further (focused) animal testing. Mathematically, a PBPK model is a group of mass balance differential equations describing the rate of change of xenobiotics and metabolites in a simulated organ­ ism that includes the basic or more detailed physiological structures (Reddy et al., 2005). Toxic effects can be studied with a linked pharmacodynamic module to examine dose– response relationships in a computer-generated environment.

73.2.2.1  Parameters Needed to Build PBPK/PD Models The procedure for developing a PBPK model is a dynamic process in which model structure can include a variety of physiological, biophysical, biochemical, and pharmaco­ dynamic parameters as illustrated in Figure 73.1. Physiological parameters refer to blood flow, compartment volume, cardiac output, and so on. Values and plausible ranges for many physiological parameters can be found in

Chapter | 73  Application of Physiologically Based Pharmacokinetic/Pharmacodynamic Modeling

Determination of model purpose

Physiological model structure

Mathematical representation

Complete literature research

Physiological parameters

Biophysical parameters

Biochemical parameters

Computer implementation and simulation of dose metrics

Model verification, validation/calibration

Pharmacodynamic parameters

Parameter value adjustment

Experimental measurements for toxicological endpoints (dose metrics)

Sensitivity, variability, and uncertainty analysis

Figure 73.1  Procedures for the construction of PBPK/PD models. Dashed arrows indicate the need for parameter value adjustment within a reasonable range.

Davies and Morris (1993), Brown et al. (1997), and U.S. EPA (2006a) and the references cited therein. Biophysical parameters mainly include partition coef­ ficients for parent compounds and metabolites. For the convenience of building PBPK models for pesticides, the partition coefficients in various tissues or organs in the rat and human have been predicted with computational mod­ els (Knaak et al., 2004, 2008, personal communication; Poulin and Krishnan, 1995a,b, 1996a,b, 1998, 1999; Poulin et al., 1999). For NMCs, partition coefficients were pre­ dicted without conside­ration of protein binding (Knaak et al., 2008). The main biochemical parameters refer to metabolic and excretionary parameters, such as Vmax and Km, when the metabolism and excretion are modeled as saturable Michaelis-Menten kinetics (Belfore, 2005; Krishnan and Anderson, 2008). Usually these metabolic parameter values come from the in vitro studies, and, more recently, they can be predicted with quantitative structureactivity relationship (QSAR) models providing modelers with at least initial values for their model construction in the expectation of saving much intensive lab work. For risk assessment purposes, AChE inhibition is the toxicological endpoint for organophosphorus insecticides

1595

and NMCs. The investigation of AChE activity in the brain or red blood cells consists of the pharmacodynamic mod­ ule in an intact PBPK/PD model. This toxicological event is modeled as a bimolecular enzyme inhibition process by NMCs (Knaak et al., 2008; Zhang et al., 2007). The most important parameter is the bimolecular enzyme inhibition rate constant (ki), which describes the inhibition capability of an inhibitor toward the enzyme. Experimentally, AChE inhibition (the ki values for 55 insecticides) has been mea­ sured in electric eel and bovine erythrocytes (Herzsprung et al., 1992). These reported ki values can be used as the initial values to start a draft model. A complete review of the ki values for organophosphorus insecticides can be found in Knaak et al. (2004) and for the NMCs in Knaak et al. (2008).

73.2.2.2  Application of the Exposure-Related Dose Estimating Model The EPA’s National Exposure Research Laboratory has developed the Exposure-Related Dose Estimating Model (ERDEM) for PBPK/PD modeling of exposure and tis­ sue dosimetry resulting from environmental chemicals (Blancato et al., 2006; U.S. EPA, 2006c). The differential mass balance equations describing the disposition and tox­ icity of chemicals in the body can be found in Blancato et al. (2006). The intact package is downloadable for free from the EPA website (U.S. EPA, 2006c). The most recent downloadable version of ERDEM is version 5.1. ERDEM exports particular model specifications, which are input by the user into a graphical user interface (GUI), into an advanced continuous simulation language (ACSL) com­ mand file to conduct a PBPK/PD model simulation. The ERDEM modeling framework has the tools to handle a wide variety of model parameters and perform stud­ ies using both simple and complex model structures. For physiological structure, a flow-limited construct is used. ERDEM is a robust modeling platform that allows spe­ cific simulation of mass transport and internal doses within the lab animal or human body. One example of a detailed PBPK model structure with full gastrointestinal (GI) com­ partments is shown in Figure 73.2, which has been used in a rat model for carbofuran (Zhang et al., 2007).

73.2.2.3  Model Calibration/Validation One important concept that is usually used by modelers is dose metrics. This refers to the target tissue dose that is closely related to ensuing adverse responses in an organism (U.S. EPA, 2006b). In the PBPK/PD modeling process, it is the point of interest that the modelers are trying to simu­ late or predict, for example, the toxic moiety concentration in the blood or brain AChE activity in the study of NMCs. With the model structure determined and metabolic reac­ tions/metabolites selected, the next step is to find the values

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Brain Static lung Bolus dose ingestion

Food flow Duodenum Small Intestine

Portal blood

Stomach

Bile flow Venous blood

Arterial blood

Colon Liver

Fat

Kidney

Fecal elimination

Urinary excretion

Slowly perfused tissue Rapidly perfused tissue

Dermis Figure 73.2  ERDEM PBPK model structure in the rat (from ������� Zhang et al., 2007; cited with permission) �������������������������������������������������� for the exposure scenario of bolus oral ingestion included a full gastrointestinal compartment and enterohepatic circulation of glucuronic acid conjugates.

for various model parameters. It is always a challenge to find the experimental evidence, or, in a better situation, there are some experimentally measured values available, but they cannot be used directly in the model. A complete literature search is necessary to reveal the variation of parameter val­ ues. Moreover, the diverse experimental data sets need to be identified for possible use in model calibration and validation. With the initial values and experimental pharmacokinetic or pharmacodynamic data available, the draft model will run against the dose metrics reported in the literature. It is worth mentioning that sometimes raw data values are not directly reported in the references and only plots and curves are avail­ able. Special digitizing software, such as DigitizeIt (ShareIt!, Cologne, Germany), is needed to digitize the plots into numerical time-course values. In other situations, the math­ ematical units need to be converted from the in vitro data into in vivo data that can be further used in the model construc­ tion. In vitro metabolic studies are one of the typical examples for which the unit conversion needs to be considered. Based on the results of the initial fitting, the preliminary simulation

may not be satisfactory; consequently, some model parameter values may be deemed necessary for consideration of adjust­ ment within a reasonable range. This adjustment of parameter values is referred to as the process of parameter value optimi­ zation. The optimization process is usually iterative and the process is repeated until the best available fit to the experi­ mental measurements is achieved by either visual inspection or statistical analysis. The procedures used to derive a final­ ized PBPK model are summarized in Figure 73.1. Under most situations, the model-predicted dose metrics cannot fit to all of the data sets from different studies closely and concurrently, or the dose metrics may just fit a part of the time history data. Therefore, not surprisingly, as more experimental dose metrics are included in the simulation, it is more likely that discrepancies will be seen since uncer­ tainties exist both in the model itself as well as in the experi­ mental data sets from different studies. However, a model constructed with as many diversified data sets as possible will be more dependable and will have less uncertainty when extrapolations are performed than a model constructed with

Chapter | 73  Application of Physiologically Based Pharmacokinetic/Pharmacodynamic Modeling

fewer experimental data sets. A PBPK/PD model for carbo­ furan in the Sprague-Dawley (SD) rat was built with diverse data sets including different exposure routes and doses (Zhang et al., 2007). For the purpose of evaluating good­ ness of fit, traditional statistical procedures aimed at analyz­ ing whether the underlying distributions of the two data sets are similar or not, such as t-, Mann-Whitney, two-sample 2, and two sample Kolmogrov tests, are not applicable (U.S. EPA, 2006a). Due to the difficulty of regular statisti­ cal approaches, so far the most convenient and still accept­ able way to judge the goodness of fit is by visual inspection to evaluate how close the simulated curves are to the data points (U.S. EPA, 2006a). Readers are reminded that visual judgment is affected by how the simulated curves are pre­ sented. For example, a logarithmic scale of the dose metrics will visually reduce the discrepancy between the simulated curve and the experimental data points. Traditional one- and two-compartment kinetic mod­ els are useful in describing the kinetics of a chemical for any available data set, but these models cannot be used for extrapolating beyond the data used to develop the model (U.S. EPA, 2006a). A good and useful PBPK model is one that, theoretically, can simulate an independent data set from a different group of researchers or any independent exposure scenarios that have never been used in the model param­eter optimization process. The models that are useful in risk assessment should be able to concurrently integrate diverse pharmacokinetic data under various exposure routes and make predictions on tissue dosimetry or toxicity beyond the data sets used for model optimization. Since these data sets are not used for model parameter optimization, simu­ lated results running against these data will provide strong evidence of the predictability of the model. Such models are valuable for risk assessment since they are capable of predicting the in vivo pharmacokinetic profiles at very low exposure levels that are applicable to human environmental exposures (U.S. EPA, 2006a). Whether a constructed model is useful or not is also decided by general model behavior. A dependable model should take into account the important pharmacokinetic characteristics of a chemical, such as the overall half-life, bioavailability, percentage of dose excreted through kidneys and bile, and dose eliminated in feces. Such an effort had been implemented in the carbofuran model construction for the SD rat (Zhang et al., 2007)

73.2.2.4  Postmodel Construction Analysis It should be noted that the calibrated and validated PBPK model represents the average individual in laboratory animal or human populations. The model parameter val­ ues stands for the population mean and do not reflect population variability. Thus, compared with the prob­ abilistic approach, the constructed model is deterministic. Sometimes, it is called a baseline model by some model­ ers (Zhang et al., 2007). After a model has been completed

1597

with a fit to all the available experimental data as satisfacto­ rily as possible, modelers may further perform extra analy­ ses on the finished model. These analyses are generalized here as postmodel construction analysis, which includes sensitivity analysis for input model parameters, variabil­ ity analysis for model output predictions, and uncertainty analysis for the model itself. The concepts and method­ ologies of these analyses will be briefly discussed here. Readers can find more discussion on these topics in Chiu et al. (2007). Sensitivity analysis is meant to find out how model input parameter values influence the estimates of the dose metrics or model predictions. When only one parame­ ter varies at a time, it is named as local (or functional) sen­ sitivity analysis, while global sensitivity analysis refers to all parameters being varied simultaneously. Parameters that have the greatest impact on the model outputs of interest can then be studied in future efforts to reduce the uncer­ tainty of these key players. For detailed methodology to perform sensitivity analysis, readers are referred to U.S. EPA (2006a) for principles and Zhang et al. (2007) for an example of sensitivity analysis in which only one param­ eter value was perturbed at one time while all the others were held constant (local sensitivity analysis). The next step is variability analysis, which is meant to evaluate the impact of the variability of model input parameters on the variability of the model output (dose metrics). The model calibrated with experimental data rep­ resents only the average individual of an animal or human population. By considering the fluctuation of input model parameter values in a population, the population range or variability of the dose metrics needs to be known. For this purpose, Monte Carlo sampling techniques based on the distributions of input parameters have been used. Readers can go to U.S. EPA (2006a) for further details and can find the examples in the cited references therein. In performing the Monte Carlo simulations, it should be noted that depen­ dency among some model parameters should be considered whenever a parameter value is perturbed in sensitivity and variability analysis. The rest of the dependent parameters will also need to change their values in order to keep the mass balance. For example, the fractional blood flows and compartment volume should be summed to 100%. No mat­ ter what value will be selected to perturb for any param­ eter, the rest of the parameter values need to be updated in order to keep the total 100%. An example with such a consideration can be found in Zhang et al. (2007). Uncertainty analysis for PBPK models evalu­ ates the impact of lack of information about either the numer­ical values of model parameters or model struc­ ture on model-predicted dose metrics (U.S. EPA, 2006a). Uncertainty analysis is particularly useful when a PBPK model does not simulate the experimental data well enough. Quantitative uncertainty analyses can be per­ formed by using a traditional Monte Carlo approach, a Bayesian Markov chain Monte Carlo analysis, a stochastic

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Synthesis of new AChE ks k+1 [ AChE – I ] AChE + I k-1

kc

kr H2O

kd

Leaving g roup

ki

CO2

O II [ AChE -C-NHCH3 ]

Deg radation

(A)

te

lea Re

d se

li bo ta e M

1 Kr

rb Ca

Synthesis of new AChE Ca

AChE + NH2CH3

rba ma te

Released Metabolite 1 Kr

Kr2

Ki2

Carbamate 2+

Free AChE

te1 ma a b ar

Carbamate2-AChE Kr2

Degradation of AChE

Ki

Re lea

N

Kd

+

C

E

1

Ki

Ks

N+

Ch -A

1 a te am

Kr

N Ca rb

(B)

se dM

Kr N am a te

et a bo li

te

N-

AC h

E

Figure 73.3  Pharmacodynamic model for AChE inhibition by carbofuran and its metabolite, 3-hydroxycarbofuran (A) (from Zhang et al., 2007; cited with permission), and the model in the cumulative PBPK/PD model for NMCs (B). There is no interaction among the members of NMCs, and nei­ ther is there competition from the NMCs with AChE. I, inhibitor; ks, resynthesis rate constant of enzyme; kd, degradation rate constant of the inhibited enzyme; kr, regeneration rate constant of the inhibited enzyme; ki, enzyme inhibition rate constant; k1, k1, and kc were not modeled.

response surface method, and a fuzzy simulation approach (U.S. EPA, 2006a, and the references therein). In gen­ eral, sensitivity, variability, and uncertainty analyses can improve the credibility of PBPK models and can also help prioritize research needs to reduce uncertainties in the developed model used in risk assessment. So far such analyses may not be required for all PBPK models for the purpose of risk assessment (U.S. EPA, 2006a).

73.3  Application of the constructed PBPK/PD model 73.3.1  Toxicity Study of Carbofuran The PBPK/PD model for carbofuran has been constructed for the SD rat (Zhang et al., 2007), and the corresponding human model was derived by replacing the rat physiological structure with that of the human and compensating for the lower oxidation of carbofuran. The physiological struc­ ture included arterial blood, brain, skin, fat, GI tract, kidney, liver, rapidly perfused tissue, slowly perfused tissue, static lung, portal blood, and venous blood. Nonperfused tis­ sue was not modeled as a separate compartment. A full GI

compartment model including stomach, duodenum, lower small intestine, and colon was utilized to better describe carbofuran GI absorption, biliary circulation, and fecal elim­ ination with considerable elaboration (Figure 73.2). Both the parent and its oxidized metabolite, 3-hydroxy carbofuran, are AChE inhibitors (Herzsprung et al., 1992; Knaak et al., 2008). With this consideration, a complete metabolic path­ way for carbofuran was incorporated into the model for SD rats (Zhang et al., 2007). The AChE inhibition process was modeled as bimolecular inhibition process (Hetnarski and O’Brien, 1975) using a bimolecular enzyme inhibition rate constant (ki) as indicated in Figure 73.3A. Once going through all the necessary procedures men­ tioned previously, the constructed PBPK/PD model can be further used for various applications. With the carbo­ furan model in the SD rat available (Zhang et al., 2007), the dose–response relationship by oral exposure in the rat was studied by simulating a series of exposure doses using ERDEM. These results are not published anywhere else and they are presented here only for the purpose of dem­ onstrating the methodology. The simulations were based on the scenario reported by Ferguson et al. (1984), which was used as the major data set for model construction. The NOAEL was then selected from the dose–response curve

Chapter | 73  Application of Physiologically Based Pharmacokinetic/Pharmacodynamic Modeling

1599

100 100

80

80

20

60 40

1000

5000

20

4000

(A)

4000 5000

400

400 300 200 100

e Tim

in)

Tim 200 e( m/ n)

(m

0

2000

300

(B)

1000

se

3000

3000

) g/kg

e (µ Dos

2000

(µg /kg )

(%) AChE

40

Do

(%) AChE

60

100 0

Figure 73.4  Three-dimensional plotting of the time course of AChE activity in the blood simulated by the model at different dosages administered orally to the rat. A front view; B back view. The dots represent the experimental data reported by Ferguson et al. (1984).

100

80 AChE activity (%)

by targeting a 10% AChE inhibition in the blood (BMD10). A dose range from 1 to 5000 g/kg was simulated at the needed increment from 5 to 1000 g/kg by oral exposure to the SD rat. Time, carbofuran dosage, and AChE activity (% of control) were plotted in a 3-D curve (Figure 73.4). Without considering the time to reach the minimum activ­ ity (or maximum inhibition), the minimum AChE activity under each dosage was plotted against the exposed dos­ ages. This minimum AChE curve is actually the bottom (the blade) of the 3-D valley-shaped sheet. For convenience of presentation, this bottom curve was plotted in a logarith­ mic dosage-minimum activity shown as an S-shaped curve in Figure 73.5. The toxicological endpoint, such as NOAEL or BMD, could be determined from this curve. As a result, a NOAEL or BMD10 value (central estimation) of 10 g/kg was determined by targeting 90% of control AChE activity (10% inhibition) in red blood cells (RBCs) as the endpoint. This BMD10 value was actually the same as what was used for the critical toxicity endpoint in the risk characterization for carbofuran in the California Environmental Protection Agency (Cal/EPA) report (Cal/EPA, 2006), in which an acute regulatory LED05 value (the lower bound on the effective dose at the 95% confidence limit) of 0.01 mg/kg was assigned. The advantage that can be seen here is that the application of the constructed PBPK/PD model enables the dose–response study and the toxicological endpoints to be estimated in silico, something that cannot be easily achieved by regular bench work. But readers are reminded again that only when a well-calibrated and validated PBPK/ PD model is constructed would these computerized study findings be useful.

60

40

20

0 1

5

10

50 100 500 Log Dosage (µg/kg)

5000

Figure 73.5  Minimum AChE activity (% of control) in the blood plot­ ted with logarithmic change of different dosages after carbofuran oral exposure to the rat.

73.3.2  Construction and Application of a Cumulative PBPK/PD Model of Three NMCs When multiple chemicals of a CMG are studied for cumula­ tive toxicity, a model will be developed that includes all par­ ent chemicals plus corresponding metabolites and associated metabolic pathways. Furthermore, when multiple exposure pathways are to be examined, the cumulative model should also include multiple routes of entry. Therefore, a ­cumulative

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PBPK/PD model is one that simulates concurrent mul­ tiple chemical exposures and includes all related routes of entry. Such a PBPK/PD model for a mixture has no formal nomenclature. The term cumulative model is used here for convenience. A cumulative model is built by combining the individual PBPK/PD models for each chemical together into one system. These individual models must have gone through the model construction procedures including model param­ eter optimization, model calibration/validation, and post­ model construction analysis in addition to quality assurance so that confidence in using these models can be established. A PBPK/PD model for the binary mixture of chlor­ pyrifos and diazinon in the rat has been built (Timchalk and Poet, 2008). Similarly, a cumulative model for three NMC insecticides (carbofuran, aldicarb, and carbaryl) has been constructed (Zhang et al., 2008, personal communi­ cation). For simplification, the completed individual model for each exposed chemical may not necessarily be entirely included into the cumulative model. For example, metabol­ ites that are not used as biomarkers can be removed from the individual mode. Only those aspects that are related to the dose metrics of interest or toxicities are needed while the same physiological structures are kept. The sim­ plified individual models have a reduced number of param­ eters, which lessens the burden of computer simulation for the assembled cumulative model. In the cumulative model of three NMCs, the interactions from the mixture of AChE inhibitors, for example, the competitive interaction from all enzyme inhibitors, are not considered in the model. This individual action toward AChE activity was shown in Figure 73.3B. However, if there is strong evidence showing a competitive interaction, such as competition for the meta­ bolic enzymes (e.g., cytochrome P450) among metabolic substrates, then the competition for the enzymes by various substrates should be considered. If the exposure level is

very low such that the tissue concentration of one substrate is far less than its corresponding Ki value (I    Ki), then the competition for the metabolic enzyme by this substrate would be ignored. This rule can be verified in the equation of competitive action toward the metabolic enzyme shown in Eq. (2), when the mechanism is described by the com­ petitive Michaelis-Menten kinetics. V  Vmax

[S]  [ I ] K m 1    [ S ]  K  i



(2)

where V is the reaction rate; Vmax is the maximum reac­ tion rate for one substrate; Km is its Michaelis constant; [S] is its concentration; [I] is the concentration of the competi­ tor; and Ki is the Michaelis constant of the competitor.

73.3.2.1  Exposure Assessment as Input for the Cumulative Model To use the cumulative model in risk assessment forwardly (compared to using it for exposure reconstruction), the model required exposure input that describes the potential chemical exposure at the time of contact with the human body. The methodology is briefly summarized in Figure 73.6. For pes­ ticide exposure, the major sources include food and drinking water ingestion, contact with residues in breathable air, and dermal contact with surfaces (U.S. EPA, 2002a). The cumulative PBPK/PD model simulated aggregate exposure to three NMC insecticides: (1) carbaryl from food intake, (2) aldicarb in drinking water, and (3) carbofuran in food via dietary ingestion. Exposure inputs for the three NMCs via the oral pathway were taken from outputs of the Stochastic Human Exposure Dose Simulation (SHEDS) model (Zartarian et al., 2007). SHEDS, as developed by

Aggregate exposure PBPK/PD Activity and environmental concentration distributions

lung brain kidney stomach skin

Biomarker distributions

liver

Exposure pathways Figure 73.6  Aggregate exposures are used as the exposure input for cumulative PBPK/PD models so that the dose metrics (biomarker level in this example) and their population variability can be predicted.

Chapter | 73  Application of Physiologically Based Pharmacokinetic/Pharmacodynamic Modeling

the U.S. EPA National Exposure Research Laboratory, is a probabilistic model that predicts longitudinal 1-year expos­ ure profiles for pesticide exposure assessment (U.S. EPA, 2007; Zartarian et al., 2000). The PBPK/PD model for aldicarb (Zhang et al., personal communication) was devel­ oped parallel to carbofuran and carbaryl models (Zhang et al., 2008, personal communication).

73.3.2.2  Application of the Cumulative NMC Model for Cumulative Risk Assessment In this section, the methodology of using the cumu­lative model to predict the toxicological endpoints will be demonstrated. Readers should be aware that this pilot work has not as yet been published in peer-reviewed scientific jour­ nals and is cited only to demonstrate its potential feasibility. It is recommended that readers focus on the process rather than emphasizing the results or making any conclusions. With SHEDS output available, the resulting exposure event timeline scenarios were run to predict the outcome using the cumulative NMC PBPK/PD model (Zhang et al., 2008). The simulation results were then projected to the U.S. population (Table 73.3). Distributions of AChE activities in RBCs and brain and urine biomarker concentrations were evaluated. The model prediction of the minimum AChE activities in the blood and brain were above 99.99% of control level for all age and gender groups at the 95th percentile (Table 73.3). Conceptually, model predictions of elimination are expected to agree with biomonitoring findings for similar human popu­ lations. Significant differences between model predictions and biomonitoring results might indicate model deficiencies or unaccounted pathways and routes of exposure when bio­ monitoring results are greater than modeling outcomes.

1601

To evaluate how close these predictions were to actual measurement, the modeling outcomes were compared with the National Health and Nutrition Examination Survey (NHANES) biomonitoring results reported by the U.S. Centers for Disease Control and Prevention (CDC)’s National Center for Health Statistics (CDC, 2005). NHANES assesses the exposure of the U.S. popu­ lation to environmental chemicals using biomonitoring in which chemicals or their metabolites were measured in blood and urine samples. In one example (carbaryl), the modeled cumulative elimination of 1-naphthol was sig­ nificantly lower than the measured concentrations at the 90th and 95th percentiles for groups in the national popu­ lation sorted by age and gender (Table 73.4). This dispar­ ity might be attributed to the commonality of 1-naphthol as a biomarker for other chemicals or the incomplete­ ness of the exposure that might involve additional unrep­ resented pathways and routes in the model simulation for carbaryl. This is one of the weaknesses that the read­ ers should be aware of in using the PBPK method for CRA. The reliability of using the PBPK model approach will be limited if the exposure inputs to the model are only partial. Related to carbofuran, carbofuran phenol was mea­ sured as the biomarker of carbofuran exposure. The model­simulated biomarker levels were in trace levels and close to what had been detected (Zhang et al., 2008). Even though this comparison did not use any statistical procedure due to the lack of information on both sampled populations, it shows that the application of the cumulative model demon­ strates a plausible method for in silico human population risk assessment. However, results from NHANES will be the final gold standard for comparison.

Table 73.3  Cumulative Model Simulation for the Minimum AChE Activity (% of Control) in the Blood and Brain Using the SHEDS Exposure Data for 1-Day and 7-Day Longitudinal Study Simulation duration 1 daya

7 daysb

Age group (years) 6–59

6–14

15–59

a

n

19,724

2,602

7,589

Compartment

Selected Percentiles

Mean 25th

50th

75th

90th

95th

Brain

100

100

99.9998

99.9993

99.9966

99.9931

Venous blood

100

100

99.9999

99.9996

99.9981

99.9964

Brain

99.9966

100

99.9999

99.9986

99.9953

99.9911

Venous blood

99.9982

100

99.9999

99.9992

99.9973

99.9954

Brain

99.9974

100

99.9998

99.9992

99.9971

99.9936

Venous blood

99.9983

100

99.9999

99.9996

99.9983

99.9966

SHEDS database provided 2 nonconsecutive days of oral exposure data. They were treated as 2 days in terms of person-days. The distribution was for SHEDS sample population instead of projected U.S. population. b The distribution is based on the projected 1995 U.S. population.

Hayes’ Handbook of Pesticide Toxicology

1602

Table 73.4  Cumulative Model Simulation for the Urinary Biomarker (1-Naphthol) Concentration for Carbaryl Using the SHEDS Exposure Data for 1-Day and 7-Day Longitudinal Studies and Comparison with NHANES Study

Age group (years)

n

Unit of urinary concentrationa

Mean

Selected Percentiles 25th

50th

75th

90th

95th

12.5

22.3

NHANES (CDC, 2005)

2,748

g/l

2b

0

1.72

4.76

Meeker et al., 2007

370

g/l

2

0

2.86

4.49

7.61

13.3

1 dayc

6–59

19,724

g/l

0.213

0

0

0

0.171

0.620

7 daysd

6–14

2,602

g/l

0.322

0

0

0

0.251

0.884

g/g Cn

0.364

0

0

0

0.266

0.948

g/l

0.202

0

0

0

0.170

0.582

g/g Cn

0.186

0

0

0

0.151

0.534

15–59

7,589

a

Urinary concentration was calculated based on the cumulative excretion normalized either by the volume or by the creatinine (Cn). Not reported. c SHEDS database provided 2 nonconsecutive days oral exposure data. They were treated as 2 days in terms of person-days. d The distribution is based on the projected 1995 U.S. population. b

73.4  Advantages and weaknesses of the cumulative PBPK/PD strategy The PBPK/PD model approach showed progress in simu­ lating the exposure scenarios in a more pharmacologically representative fashion, thus providing another choice in the analysis of the cumulative risk from random exposures. Moreover, toxicological endpoints such as AChE activity and urinary biomarker levels could be studied directly at any time without uncertainty factors. Therefore, a PBPK/ PD model approach allows the CRA to be performed with more realistic assumptions by closely simulating the expo­ sure scenarios. But in the SHEDS simulations, the cumu­ lative model for the three NMCs considered only body weight as the individual difference. Population variations such as compartment volumes, blood flows, and metabolic polymorphisms were not included (Zhang et al., 2008), even though they were technically plausible. This study demonstrated that the use of a composite PBPK/PD model linked to an exposure model, such as SHEDS, for pesticide residues in humans may provide a promising way to do in silico human population risk assessment. The method of using PBPK/PD modeling in CRA has been regarded as one of the approaches, but only when a highly refined assessment is needed (Boobis et al., 2008). It should be noted that not all PBPK mod­ els are useful for risk assessment applications. If the ultimate goal of developing a model is for this purpose, the requirements for the quality of the model should be higher. Therefore, this data-intensive task will require

experimental ­ studies to be ­ developed specifically for modeling purposes. Experimental pharmacokinetic data targeted for laboratory animal PBPK model construction are needed. Meanwhile, a complete set of standardized pharmacokinetic data should be collected, preferably by the same group of researchers. But for human models, it is always a challenge to have any data related to the dose metrics concerned in risk assessment. Alternatively, a human model must be built by extrapolation from an ani­ mal model. Therefore, the uncertainty during the model extrapolation may be carried down to the human model. Compared with other methods, PBPK modeling provides the foundation for the study of dose metrics in a targeted human population using the stochastic technique. To do so, the model parameters need to reflect the population variability and distribution in order for the Monte Carlo sampling technique to be implemented. This information is still a challenge to all modelers. Even though the range and distribution have been determined for those physi­ ological parameters (U.S. EPA, 2006a), the parameters such as partition coefficients, metabolic parameters, and even pharmacodynamic parameters still need more inves­ tigation. Therefore, a true meaningful population model simulation will require more research. Lastly, to demon­ strate model reliability, which is especially important for regulatory decision making, formal statistical analyses are necessary, for example, uncertainty analysis for the model and Bayesian analysis for parameter optimization. However, intensive computer-based data analyses are not always possible for every risk assessor.

Chapter | 73  Application of Physiologically Based Pharmacokinetic/Pharmacodynamic Modeling

73.5  Future needs for the application of PBPK modeling   in risk assessment First of all, the use of PBPK modeling in risk assessment has been increasingly regarded as an important component of chemical risk assessment. Obviously, uncertainties still exist in the developed PBPK models in the aspects of model structure, parameter values, or even experimental data used for parameter optimization. But should the existence of these uncertainties be an obstacle for the use of the PBPK models in risk assessment? Frontier application of computational technique will serve as light guiding us walking through the “dark room” of risk assessment. “Better judgments are made as result of the light, even a dim one” (Blancato, 2009). More investigations aiming to reduce the uncertainties are needed. Second, the model structures and the parameters used to describe the absorption, distribution, metabolism, excretion, and toxicity (ADMET) need international har­ monization so that models from different researchers can be convertible or even “cloned” into another system. Different modelers may use different model parameters to describe the same biological process. For example, to describe GI absorption, the absorption rate constant (Ka, h1) has been used for chlorpyrifos (Timchalk et al., 2002) and carbo­ furan (Zhang et al., 2007), while the oral absorption fraction (fabsoral) was used in a malathion model (Bouchard et al., 2003). A standardized or recommended set of model param­ eters needs to be available for use by modelers. In addition, there needs to be an effort to perform and standardize the measurement of parameter values. Thirdly, good modeling practice (GMP) for PBPK model development, character­ ization, documentation, and evaluation has been proposed (Loizou et al., 2008). The development and implementation of GMP for PBPK modeling will increase the transparency of model development and model documentation. These efforts will make the work of quality assurance and quality control more efficient and can increase the credibility of a constructed model so that the developed PBPK model can be used for risk assessment with more confidence.

Conclusion CRA is needed when exposure to multiple chemicals exert­ ing their toxicity in the same mechanism is encountered. Several methods for CRA have been used, including the toxicological index method, margin of exposure method, RPF method, and PBPK/PD modeling method. In the RPF method, an index chemical is chosen from the CMG and the exposure dose will be summed by each of the con­ verted equivalent dose of the index chemical based on each chemical’s RPF. The PBPK/PD modeling approach can simulate the exposure in a more pharmacokinetic fashion and make predictions on the toxicological endpoints. But

1603

models built for such a purpose need to be of higher qual­ ity and must be constructed with diversified experimental data, model calibration/validation, and uncertainty analy­ sis. Therefore, more research efforts are needed to build more dependable PBPK/PD models for their use in CRA.

Acknowledgments Although this work was reviewed by the U.S. EPA and approved for publication, it does not necessarily reflect official Agency policy or the official views of General Dynamics Information Technology (GDIT). Mention of trade names or commercial products does not constitute endorsement or recommendation for use. The authors gratefully acknowledge Jerry J. Lorenz, Andy M. Tsang, and Lynda S. Harrison of GDIT for their great help in the preparation of this manuscript. In mem­ ory, the authors also wish to acknowledge Dr. Frederick W. Power, who was one of the primary founders and the chief architect of the ERDEM platform, who tirelessly pursued the development and enhancement of ERDEM’s scope and efficacy until his last days. Our exposure-dose assessment work using ERDEM would not have been possible without his great contribution.

References Baron, R. L. (1991). Carbamate insecticides. In “Handbook of Pesticide Toxicology” (W. J. Hayes Jr. and E. R. Laws Jr., eds.), pp. 1125– 1189. Academic Press, New York. Belfore, C. J. (2005). Pesticides and persistent organic pollutants (POPs). In “Physiologically Based Pharmacokinetic Modeling: Science Applications” (M. Reddy, R. S. Yang, M. E. Andersen, and H. J. Clewell, eds.), pp. 172–182. John Wiley & Sons Inc., Hoboken, New Jersey. Blancato, J. N. (2009). Computational toxicology in cancer risk assessment cancer risk assessment. In “Chemical Carcinogenesis from Biology to Standards Quantification” (C.-H. Hsu and T. Stedeford, eds.). McGraw-Hill, Hoboken, New Jersey. Blancato, J. N., Power, F. W., Brown, R. N., and Dary, C. C. (2006). Exposure Related Dose Estimating Model (ERDEM): A Physiologically Based Pharmacokinetic and Pharmacodynamic (PBPK/PD) Model for Assessing Human Exposure and Risk. U.S. Environmental Protection Agency, Washington, DC. EPA/600/R-06/061 (NTIS PB2006-114712). Available at: http://epa.gov/heasd/products/erdem/237edrb05-Report.pdf. Bouchard, M., Gosselin, N. H., Brunet, R. C., Samuel, O., Dumoulin, M. J., and Carrier, G. (2003). A toxicokinetic model of malathion and its metabolites as a tool to assess human exposure and risk through mea­ surements of urinary biomarkers. Toxicol. Sci. 73(1), 182–194. Available at: http://toxsci.oxfordjournals.org/cgi/content/abstract/73/1/182. Boobis, A. R., Ossendorp, B. C., Banasiak, U., Hamey, P. Y., Sebestyen, I., and Moretto, A. (2008). Cumulative Risk Assessment of Pesticide Residues in Food. Toxicol. Lett. 15,180(2), 137–150. Available at: http://www.ncbi.nlm.nih.gov/pubmed/18585444. Brown, R. P., Delp, M. D., Lindstedt, S. L., Rhomberg, L. R., and Beliles, R. P. (1997). Physiological parameter values for physiologically

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based pharmacokinetic models. Toxicol. Ind. Health 13, 407–484. Available at: http://www.ncbi.nlm.nih.gov/pubmed/9249929. Cal/EPA (2006). Risk Characterization Document: Carbofuran. Department of Pesticide Regulation, California Environmental Protection Agency. Available at: http://www.cdpr.ca.gov/docs/risk/rcd/carbofuran.pdf. CDC (2005). National Health and Nutrition Examination Survey: Third National Report on Human Exposure to Environmental Chemicals. Department of Health and Human Services, Centers for Disease Control and Prevention. Available at: http://www.cdc.gov/exposurereport/pdf/ thirdreport.pdf. Chiu, W. A., Barton, H. A., DeWoskin, R. S., Schlosser, P., Thompson, C. M., Sonawane, B., Lipscomp, J. C., and Krishnan, K. (2007). Evaluation of physiologically based pharmacokinetic models for use in risk assess­ ment. J. Appl. Toxicol. 27, 218–237. Conolly, R., Wang, J., and Tan, Y. (2005). Physiological based phar­ macokinetic (PBPK) modeling as an alternative to relative potency factors (RPFs) in Cumulative risk assessment: an example with ace­ tylcholinesterase (AChE) inhibiting pesticides. Toxicol. Sci. 84(S-1), 7 (Abstract 34). Davies, B., and Morris, T. (1993). Physiological parameters in laboratory animals and humans. Pharm. Res. 10(7), 1093–1095. Available at: http://www.springerlink.com/content/j302232361082w14/ Ecobichon, D. J. (1991). Toxic Effects of Pesticides Chapter 18. In “Casarett and Doull’s Toxicology: The Basic Science of Poisons” (M. O. Amdur, J. Doull, and C. D. Klaassen, eds.) 4th ed., pp. 565– 623. Pergamon Press, New York, NY. Ferguson, P. W., Dey, M. S., Jewell, S. A., and Krieger, R. I. (1984). Carbofuran metabolism and toxicity in the rat. Fund Appl. Toxicol. 4, 14–21. Herzsprung, P., Weil, L., and Niessner, R. (1992). Measurement of bimo­ lecular rate constants Ki of the cholinesterase inactivation reaction by 55 insecticides and of the influence of various pyridimium oximes on Ki. Int. J. Environ. Anal. Chem. 47, 181–200. Available at: http://cat .inist.fr/?aModele  afficheN&cpsidt  5397833. Hetnarski, B., and O’Brien, R. D. (1975). Electron-donor and affinity constants and their application to the inhibition of acetylcholinester­ ase by Carbamates. J. Agric. Food Chem. 23(4), 709–713. Knaak, J. B., Dary, C. C., Power, F., Thompson, C. B., and Blancato, J. N. (2004). Physicochemical and biological data for the ­development of predictive organophosphorus pesticide QSARs and PBPK/PD models for human risk assessment. Crit. Rev. Toxicol. 34(2), 143–207. Knaak, J. B., Dary, C. C., Okino, M. S., Power, F. W., Zhang, X., Thompson, C. B., Tornero-Velez, R., and Blancato, J. N. (2008). Parameters for carbamate pesticide QSAR and PBPK/PD models for human risk assessment. Rev. Environ. Contam. Toxicol. 193, 53–210. Krishnan, K., and Anderson, M. E. (2008). Chapter 5. Physiologically based pharmacokinetic and toxicokinetic models. In “Principles and Methods of Toxicology” (A. W. Hayes ed.) 5th ed., pp. 239–242. Taylor & Francis, Boca Raton, FL. Kuhr, R. J., and Dorough, H. W. (1976). “Carbamate insecticides: Chemistry, biochemistry and toxicology,” pp. 41–66. CRC Press, Cleveland, OH. Loizou, G., Spendiff, M., Barton, H. A., Bessems, J., Bois, F. Y., d’Yvoire, M. B., Buist, H., Clewell, H. J. 3rd, Meek, B., GundertRemy, U., Goerlitz, G., and Schmitt, W. (2008). Development of good modeling practice for physiologically based pharmacokinetic models for use in risk assessment: The first steps. Regul. Toxicol. Pharmacol. 50(3), 400–411. Available at: http://www.ncbi.nlm.nih .gov/pubmed/18331772. Lowit, A., Dary, C., Power, F., Blancato, J., Setzer, R. W., Conolly, R., and Seaton, M. (2004). Physiologically-Based Pharmacokinetic/

Hayes’ Handbook of Pesticide Toxicology

Pharmacodynamic Modeling and Cumulative Risk Assessment: Case Study for the N-Methyl Carbamate Pesticides. Available at: http:// www.epa.gov/nerl/news/forum2004/lowit_web.pdf. Matsumura, F. (1985). “Toxicology of Insecticides,” pp. 161–177 2nd ed.. Plenum Press, New York, NY. Meeker, J. D., Barr, D. B., Serdar, B., Rappaport, S. M., and Hauser, R. (2007). Utility of Urinary 1-naphthol and 2-naphthol levels to assess environmental carbaryl and naphthalene exposure in an epidemiology study. J. Expo. Sci. Environ. Epidemiol. 17(4), 314–320. O’Brien, R. D. (1967). “Insecticides, Action and Metabolism,” pp. 86–97. Academic Press, New York, NY. Padilla, S., Setzer, W., Marshall, R. S., Hunter, D., Phillips, P., McDaniel, K., Moser, and V. C., Lowit, A. (2006). A Dose-Response Study of the Toxicity of a Mixture of 7 N-Methyl Carbamates in Adult, Male Rats. Presented at the 45th Annual Meeting of the Society of Toxciology; San Diego, California. Padilla, S., Marshall, R. S., Hunter, D. L., and Lowit, A. (2007). Time course of cholinesterase inhibition in adult rats treated acutely with carbaryl, carbofuran, formetanate, methomyl, methiocarb, oxamyl or propoxur. Toxicol. Appl. Pharmacol. 219, 202–209. Available at: http://cat.inist.fr/?aModele  afficheN&cpsidt  18632171. Poulin, P., and Krishnan, K. (1995a). A biologically-based algorithm for predicting human tissue:blood partition coefficients of organic chem­ icals. Hum. Exp. Toxicol. 14, 273–280. Poulin, P., and Krishnan, K. (1995b). A tissue composition-based algo­ rithm for predicting tissue: Air partition coefficients of organic chem­ icals. Toxicol. Appl. Pharmacol. 136, 126–130. Poulin, P., and Krishnan, K. (1996a). A mechanistic algorithm for pre­ dicting blood: Air partition coefficients of organic chemicals with the consideration of reversible binding in hemoglobin. Toxicol. Appl. Pharmacol. 136, 131–137. Poulin, P., and Krishnan, K. (1996b). Molecular structure-based predic­ tion of the partition coefficients of organic chemicals for physiologi­ cal pharmacokinetic models. Toxicol. Methods 6, 117–137. Poulin, P., and Krishnan, K. (1998). A quantitative structure-toxicokinetic relationship model for highly metabolised chemicals. Altern. Lab. Anim. 26, 45–59. Poulin, P., and Krishnan, K. (1999). Molecular structure-based prediction of the toxicokinetics of inhaled vapors in humans. Intl. J. of Toxicol. 18, 7–18. Poulin, P., Beliveau, M., and Krishnan, K. (1999). Mechanistic animalreplacement approaches for predicting pharmacokinetics of organic chemicals. In “Toxicity Assessment Alternatives: Methods, Issues, Opportunities” (H. Salem and S. A. Katz, eds.), pp. 115–139. Humana Press Inc., Totowa, New Jersey. Reddy, M., Yang, R. S., Andersen, M. E., and Clewell, H. J. (2005). “Physiologically Based Pharmacokinetic Modeling: Science Applications,������������������������������������������������������������������������� ”������������������������������������������������������������������������ John Wiley & Sons Inc., Hoboken, New Jersey. Timchalk, C., Nolan, R. J., Mendrala, A. L., Dittenber, D. A., Brzak, K. A., and Mattsson, J. L. (2002). A physiologically based pharmacokinetic and pharmacodynamic (PBPK/PD) model for the organophosphate insecti­ cide chlorpyrifos in rats and humans. Toxicol. Sci. 66, 34–53. Available at: http://toxsci.oxfordjournals.org/cgi/content/abstract/66/1/34. Timchalk, C., and Poet, T. S. (2008). Development of a physiologically based pharmacokinetic and pharmacodynamic model to determine dosimetry and cholinesterase inhibition for a binary mixture of chlor­ pyrifos and diazinon in the rat. Neurotoxicology 29(3), 428–443. U.S. EPA (1999). Significant Guidance Document: Guidance for Identifying Pesticide Chemicals and Other Substances That Have a Common Mechanism of Toxicity. U.S. Environmental Protection

Chapter | 73  Application of Physiologically Based Pharmacokinetic/Pharmacodynamic Modeling

Agency, Office of Pesticide Programs, Washington, DC. EPA-HQOPP-2007-0793-0001. Posted August 14, 2007. Available at: http:// www.regulations.gov/fdmspublic/component/main?mainDocket Detail&dEPA-HQ-OPP-2007-0793. U.S. EPA (2001). Significant Guidance Document: General Principles for Performing Aggregate Exposure and Risk Assessments. U.S. Environmental Protection Agency, Office of Pesticide Programs, Washington, DC. EPA-HQ-OPP-2007-0792.0001. Date posted: August 14, 2007. Available at: http://www.regulations.gov/fdmspublic/component/ main?mainDocketDetail&dEPA-HQ-OPP-2007-0792. U.S. EPA (2002a). Significant Guidance Document: Guidance on Cumulative Risk Assessment of Pesticide Chemicals that have a Common Mechanism of Toxicity. U.S. Environmental Protection Agency, Office of Pesticide Programs, Washington, DC. EPA-HQOPP-2007-0797.0001. Date posted: August 14, 2007. Available at: http://www.regulations.gov/fdmspublic/component/main?mainDoc ketDetail&dEPA-HQ-OPP-2007-0797. U.S. EPA (2002b). Cumulative Risk Assessment: Developing the MethodsAvailable Papers and Where They May be Located. U.S. Environ­ mental Protection Agency, Office of Pesticide Programs, Washington, DC. Available at: http://www.epa.gov/pesticides/cumulative/Cum_Risk_ AssessmentDTM.htm. U.S. EPA (2006a). Approaches for the Application of Physiologically Based Pharmacokinetic (PBPK) Models and Supporting Data in Risk Assessment (Final Report). U.S. Environmental Protection Agency, Washington, DC. EPA/600/R-05/043F. Federal Register: September 22, 2006. Volume 71, Number 184, Notices: Page 55469-55470. Available at: http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid157668. U.S. EPA (2006b). Organophosphorus Pesticides (OP) Cumulative Assessment-2006 Update. U.S. Environmental Protection Agency, Office of Pesticide Programs, Washington, DC. July 31, 2006. Available at: http://www.regulations.gov/fdmspublic/component/ main?mainDocketDetail&dEPA-HQ-OPP-2006-0618.

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U.S. EPA (2006c). Exposure Related Dose Estimating Model (ERDEM). U.S. Environmental Protection Agency, Office of Research and Development. Downloadable version available at: http://www.epa. gov/heasd/products/erdem/erdem.htm. U.S. EPA (2007). N-Methyl Carbamate Cumulative Risk Assessment. U.S. Environmental Protection Agency, Office of Pesticide Programs, Washington, DC. EPA-HQ-OPP-2007-0935-0003. Available at: http:// www.regulations.gov/fdmspublic/component/main?main  Documen tDetail&o  09000064802a02db. Zartarian, V. G., Ozkaynak, H., Burke, J. M., Zufall, M. J., Rigas, M. L., and Furtaw, E. J. Jr. (2000). A modeling framework for estimating children’s residential exposure and dose to chlorpyrifos via dermal residue contact and non-dietary ingestion. Environ. Health Perspect. 108(6), 505–514. Available at: http://www.pubmedcentral.nih.gov/ articlerender.fcgi?artid  1638152. Zartarian, V., Glen, G., Smith, L., and Xue, J. (2007). SHEDSMultimedia Model version 3 Technical Manual (Draft June 14, 2007). Pre-dissemination peer review copy at: http://www.epa. gov/scipoly/sap/meetings/2007/august/sheds_techmanual_06_14.pdf. Zhang, X., Tsang, A. M., Okino, M. S., Power, F. W., Knaak, J. B., Harrison, L. S., and Dary, C. C. (2007). A physiologically based pharmacokinetic/pharmacodynamic model for carbofuran in SpragueDawley rats using the exposure-related dose estimating model. Toxicol. Sci. 100(2), 345–359. Available at: http://toxsci.oxfordjourn­ als.org/cgi/content/full/100/2/345. Zhang, X., Okino, M. S., Knaak, J. B., Tsang, A. M., Xue, J., Harrison, L. S., Heravi, N. E., Gerlach, R.W., Johnson, J.C., Tornero-Velez, R., Zartarian, V. G., and Dary, C. C. (2008). Cumulative Risk Estimation Using Exposure and Dose Models to Three N-Methyl Carbamates: Carbaryl, Aldicarb, and Carbofuran. 47th Annual Meeting and ToxExpo, Seattle, Washington, March 16-20, 2008. Abstract 269 available at: http://www.toxicology.org/ai/pub/Tox/2008Tox.pdf.

Chapter 74

Toxicological Profile of Carbaryl Ann M. Blacker,1 Curt Lunchick,1 Dominique Lasserre-Bigot,2 Virginie Payraudeau2 and Mike E. Krolski3 1

Bayer CropScience, Research Triangle Park, North Carolina Bayer CropScience, Sophia Antipolis, France 3 Bayer CropScience, Stilwell, Kansas 2

74.1  Introduction Carbaryl is a carbamate pesticide that was first registered for use in the United States on cotton in 1959. Today, carbaryl is a widely used broad-spectrum insecticide used in agriculture, professional turf management, ornamental production, and residential settings. Carbaryl is sold under many trade names, but the most common one is Sevin. Over the years, a large dossier of toxicity, environmental fate, residue, and monitoring data have been generated, which, coupled with practical use experience, have been used to improve application methods and to refine exposure and risk assessments that support the continued registration and safe use of this insecticide.

74.2  Description, use, and biological mode of action Technical carbaryl belongs to the N-methyl carbamate chemical family, which act via inhibition of acetylcholinesterase. The pure (technical) material is a white to light tan solid with a water solubility of approximately 40 ppm at 25°C. The empirical formula of carbaryl is C12H11NO2, with a molecular weight of 201.2. Its Chemical Abstract Service (CAS) name is 1-naphthalenyl methylcarbamate and its CAS number is 63-25-2. Carbaryl end-use products are formulated as flowable concentrates, granules, wettable powders, baits, dusts, and ready-to-use sprays. Application methods include groundboom, airblast, chemigation, aerial, drop or broadcast granular spreaders, and handheld equipment such as lowpressure handwand sprayers.

74.3  Hazard characterization Carbaryl has a very robust toxicity database including acute, chronic, developmental, reproductive, and neurotoxicity Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

studies. The toxicity data summarized in this chapter come from studies conducted during the past 20 years to support the registered and approved uses of carbaryl. A review of older toxicity data can be found in Cranmer (1986), while more recently, Gunasekara et al. (2008) reviewed the environmental fate and toxicology of carbaryl.

74.3.1  Acute Toxicity Carbaryl is moderately toxic by the oral route and exhibits a low level of toxicity following dermal or inhalation exposures. In recent acute toxicity studies, the oral LD50 in rats was 614 mg/kg [vehicle of aqueous 0.5% (w/v) carboxymethylcellulose] while the dermal LD50 in rats was greater than 5000 mg/kg (Griffon, 2001a,b). An acute inhalation (nose only) exposure study in rats with carbaryl resulted in LC50 values of greater than 4.62 mg/l in males and 2.43 mg/l in females (Wesson, 2001). Carbaryl is not a dermal or eye irritant and is not a dermal sensitizer in guinea pigs (Griffon, 2001c,d,e). Rapid and reversible cholinesterase inhibition (ChEI) has been demonstrated following acute oral exposure in rats (Brooks and Broxup, 1995a). A dose-dependent inhibition of brain and red blood cell (RBC) ChE activity was observed 30 min after oral administration of carbaryl suspended in aqueous 0.5% (w/v) carboxymethylcellulose/0.1% (w/v) Tween 80. Thereafter, the level of inhibition observed at 10 and 50 mg/kg declined slowly and returned to control levels by 24 h, but at 125 mg/kg, ChE activity remained slightly below control levels at 24 h after dosing. The results for male rats are summarized in Table 74.1 and are similar to those observed for female rats. Observed clinical signs and behavioral changes were also dose-dependent and consisted of a single observation of muzzle and urogenital staining in one male at 10 mg/kg, and tremors, salivation, muzzle and urinary staining, gait alterations, decreased respiration, and decreased activity 1607

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Table 74.1  Time Course of ChEI in Male Rats Receiving a Single Oral Bolus Dose of Carbaryl Time after dosing

RBC (% control)

Brain (% control)

10 mg/kg

50 mg/kg

125 mg/kg

10 mg/kg

50 mg/kg

125 mg/kg

0.5 h

72

68

56

46

23

18

1.0 h

82

71

54

68

25

22

2.0 h

88

65

71

82

33

21

4.0 h

84

63

51

94

52

22

8.0 h

101

100

56

91

73

24

24 h

103

94

90

100

94

77

and arousal levels at 50 and 125 mg/kg. The incidence and severity of these signs decreased during the 24-h period after dosing. In another study, oral administration of carbaryl in corn oil at 30 mg/kg produced a 40% inhibition of both brain and RBC ChE activity 30 min after dosing (Padilla et al., 2007). Two hours after dosing, recovery of ChE activity to near control levels was observed for both brain and RBC, but a second phase of inhibition was noted for the RBC matrix at the 4- and 6-h time points with return to control levels by 24 h after dosing. Dermal absorption studies in rats with different formulated products show that carbaryl has a low rate of penetration. When applied to the shaved back of male rats, the mean percent of radioactivity absorbed after a 10-h exposure was 12.7 and 8.9 for a 100-fold dilution in 1.0% aqueous carboxymethyl cellulose of Sevin brand LXR Plus (a liquid formulation containing 44.1% active ingredient) and Sevin brand 80S (a dry powder formulation containing 80% active ingredient), respectively (Cheng, 1994, 1995).

74.3.2  Subacute/Subchronic Toxicity The subchronic toxicity of carbaryl has been assessed in rats, mice, and dogs. In 6- to 8-week dietary studies, a decrease of body weight and food consumption was observed at dietary levels of 3000 ppm and higher in Sprague-Dawley Crl:CD BR rats and at 4000 ppm and higher in CD-1 mice (Hamada, 1990a, b). RBC and brain ChE were decreased at 3000 ppm and higher in both rats and mice. Increased liver weight associated with centrilobular hypertrophy was observed at 6000 ppm in rats, whereas centrilobular hypertrophy without increased liver weight was noted in mice at 4000 ppm. In a 6-month dietary study in C57Bl/6 male mice, carbaryl administered at 4000 ppm resulted in slightly lower body weights and food consumption and a 14% increase in absolute liver weight without any associated histopathological changes (Chuzel, 2000a).

In a 5-week study, dogs were administered carbaryl via the diet at 0, 20, 45, and 125 ppm (Hamada, 1991). No evidence of treatment-related effects was noted for the parameters measured in this study including RBC and brain ChE. In a 1-year study conducted with dietary levels of 0, 125, 400, and 1250 ppm, clinical signs of toxicity (lacrimation, salivation, and tremors) and lower body weight gain and food consumption were observed in female dogs at 1250 ppm (Hamada, 1987). RBC ChE was significantly decreased at 400 ppm during the first half of the study and at 1250 ppm throughout the entire study. At study termination, brain ChE was 22 to 36% lower than control at 400 and 1250 ppm (both sexes) and 20% below control at 125 ppm in females. No histopathological changes related to treatment were observed. In a 4-week study, carbaryl was administered by topical application to the shaved dorsal skin of rats for 6 h/day, 5 days/week at 0, 20, 50, or 100 mg/kg/day (Austin, 2002). Lower body weight gains were observed at 100 mg/kg/day, but no treatment-related findings on mortality, clinical signs, food consumption, or dermal irritation were noted. RBC ChE, determined during each week of the study within 1 h after removal of the test material, was 10 to 20% lower than control at 50 and 100 mg/kg/day with no clear dose response. At study termination, brain ChE at these dose levels was 15 to 24% lower than control. Benchmark dose (BMD) analysis of the brain ChE data to determine a benchmark response of 10% gives a BMD10 of 49.35 mg/kg/day with a 95% lower confidence limit (BMDL10) of 30.56 mg/kg/day (US EPA, 2008).

74.3.3  Neurotoxicity There is a complete neurotoxicity database on carbaryl consisting of acute, subchronic, and developmental neurotoxicity studies conducted with Sprague-Dawley rats. Based on acute neurotoxicity testing in rats with exposure to carbaryl via oral gavage, the highest nonlethal dose was

Chapter | 74  Toxicological Profile of Carbaryl

determined to be 250 mg/kg (Brooks and Broxup, 1995b). In an acute neurotoxicity study with groups of rats dosed by oral gavage at 0, 10, 50, or 125 mg/kg, dose-related alterations in the functional observation battery (FOB) were observed 30 min after dosing (Brooks and Broxup, 1995c). FOB alterations included ataxic gait in one male and one female at 10 mg/kg and tremors; increased salivation; decreased or impaired locomotor activity, rearing, gait, arousal, and reflexes; and decreased body temperature at 50 and 125 mg/kg. In addition, increased hindlimb splay and startle response and decreased grip strength were seen at 125 mg/kg. Lower motor activity counts were observed approximately 1 h postdosing for all treated groups. No treatment-related alterations in FOB or motor activity were observed 7 or 14 days after dosing. Further, no neuropathological lesions were observed at study termination. The subchronic neurotoxicity of carbaryl was evaluated over a 13-week exposure period in rats by oral gavage at 0, 1, 10, and 30 mg/kg/day (Robinson and Broxup, 1996). Body weight and food consumption were decreased at 30 mg/kg/day throughout most of the study. Neurotoxicological evaluations conducted during weeks 4, 8, and 13 at approximately the time to peak effect, that is, between 0.5 and 1 h postdosing, revealed clinical signs and FOB changes such as increased salivation, tremors, pinpoint pupils, gait alterations, decreased rearing, reduced grip strength, as well as reduced body temperatures at 30 mg/kg/day and to a lesser extent at 10 mg/kg/day. The motor activity counts were also reduced at 30 mg/kg/day in animals at weeks 4 and 8. Inhibition of RBC and brain ChE was observed 1 h postdosing for samples collected during weeks 4, 8, and 13. At the end of the exposure period, no treatment-related neuropathological findings were observed. The no observed adverse effect level (NOAEL) was established at 1 mg/kg/day based on ChEI and alterations in the FOB. A developmental neurotoxicity study was also conducted by oral gavage at 0, 0.1, 1, and 10 mg/kg/day (Robinson and Broxup, 1997). Pregnant rats (F0) were treated from gestation day 6 (GD6) to lactation day 10 (LD10). Pups (F1) were weaned on postnatal day 21 (PND21) and observed up to PND70. Maternal performance parameters such as pregnancy rate, gestation index, sex ratio, and implantation loss were unaffected by treatment. In F0 dams at 10 mg/kg/day, tremors, reduced pupil size, and altered gait were noted during the dosing period, and RBC and brain ChE were significantly decreased at 1 h after dosing on GD20 and LD10. No neuropathological changes were observed at terminal sacrifice. For the F1 generation, no in-life phase parameters were affected including survival, body weight, clinical signs, FOB, motor activity, developmental landmarks, auditory startle, passive avoidance, and water maze measurements. Further, no changes in brain weight or neuropathology were observed at PND11 or 70. Extensive morphometric measurements showed no clear treatment-related effect at

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10 mg/kg/day. Therefore, based on observations in the FOB and on ChEI, the NOAEL for maternal toxicity was set at 1 mg/kg/day. As no treatment-related effects were observed in pups, the NOAEL for developmental neurotoxicity in the offspring was 10 mg/kg/day, the highest dose level tested. In conclusion, across all of the neurotoxicity studies, the most sensitive endpoints were observations in the FOB, reduced motor activity, and ChEI in RBC and brain. The severity and frequency of clinical signs and ChEI were dose related and decreased with time. The lowest oral NOAEL for neurotoxicity studies with carbaryl was 1 mg/kg/day.

74.3.4  Developmental and Reproductive Toxicity In recent studies, effects on fetal development were observed only in the presence of maternal toxicity. In a developmental toxicity study in Sprague Dawley Crl:CD (SD) BR rats, carbaryl was administered by gavage at 0, 1, 4, and 30 mg/ kg/day from GD6 to 20 inclusive (Repetto-Larsay, 1998). Maternal toxicity was indicated by increased salivation and approximately 30 and 15% reductions in body weight gain and food consumption, respectively, at 30 mg/kg/day. Gestational parameters were not affected. Fetal body weights were significantly reduced at 30 mg/kg/day and associated with delayed ossification of a few bones. No increased incidence of malformations was observed at any dose level. The NOAEL for both maternal and developmental toxicity was 4 mg/kg/day. In a developmental toxicity study in New Zealand White rabbits, carbaryl was administered by oral gavage at 0, 5, 50, and 150 mg/kg/day from GD6 to 29 inclusive (Tyl et al., 1999). Maternal toxicity was indicated by reduced body weight gains (25 and 53% below control for the dosing period) at 50 and 150 mg/kg/day, respectively. Maternal food consumption during the dosing period was equivalent or slightly higher than control. Gestational parameters were not affected in any treatment group. The only evidence of fetal toxicity was a 10% reduction in weight at 150 mg/kg/ day compared to control. Therefore, the NOAEL for maternal toxicity was 5 mg/kg/day and the NOAEL for developmental toxicity was 50 mg/kg/day. In a two-generation reproduction study, Sprague Dawley Crl:CD (SD) BR rats were administered carbaryl in the diet at 0, 75, 300, and 1500 ppm (Tyl et al., 2001). During premating periods, these dietary levels were equivalent, respectively, to 0, 4.67, 31.34, and 92.43 mg/kg/day for F0 males; 0, 5.56, 36.32, and 110.78 mg/kg/day for F0 females; 0, 5.79, 23.49, and 124.33 mg/kg/day for F1 males; and 0, 6.41, 26.91, and 135.54 mg/kg/day for F1 females. Parental toxicity was indicated by reduced body weight and food consumption during the premating, gestation, and lactation periods at 1500 ppm. Body weights for the F0 and Fl parental animals at 300 ppm

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were consistently lower throughout the study but were generally not statistically different from controls. No effect on reproduction, fertility, or reproductive organs was observed for any treatment group. Further, for F0 and F1 adult males, no significant differences were observed for percent motile or progressively motile sperm, epididymal sperm concentration, testicular spermatid head concentration, daily sperm production (DSP), efficiency of DSP, or percent abnormal sperm. For Fl and F2 offspring, no effects of treatment on stillbirth or live birth indices, number of live pups per litter on PND0, or sex ratio (% males) were observed. However, during the lactation period, body weights for F1 and F2 pups were significantly lower at 1500 ppm, and survival indices were decreased at 1500 ppm for F1 and F2 pups and at 300 ppm for F2 pups. The NOAEL for parental and pup toxicity was 75 ppm, and the NOAEL for reproduction and fertility was 1500 ppm. In conclusion, studies conducted on carbaryl for regulatory purposes show no evidence of teratogenicity or reproductive toxicity.

74.3.5  Genotoxicity In vitro gene mutation assays with carbaryl in both bacterial and mammalian cells are negative (Lawlor, 1989; Onfelt and Klasterska, 1984; Young, 1989). In vitro studies using Chinese hamster V79 cells indicate that carbaryl induces aneuploidy and sister chromatid exchanges (Onfelt and Klasterska, 1983, 1984). However, this cell line has been reported to contain a mutated and nonfunctional p53 protein and thus results with these cells should be interpreted with caution (Chaung et al., 1997). In the presence of metabolic activation, carbaryl did induce chromosomal aberrations in an in vitro study with Chinese hamster ovary cells (Murli, 1989), but all in vivo cytogenetics studies are negative (McEnaney, 1993; � Marshall, 1996). An in vitro DNA synthesis assay with rat primary hepatocytes is also negative (Cifone, 1991). Further, no DNA adduct formation in the liver was observed in mice administered carbaryl via the diet at 8000 ppm for 2 weeks followed by a single oral dose of C14-labeled carbaryl at 75 mg/kg (Sagelsdorff, 1994). In addition, a 6-month study in male C57BL/6 p53 heterozygous knockout mice, described in more detail in the following section, is negative (Chuzel, 2000b).

Hayes’ Handbook of Pesticide Toxicology

74.3.6.1  Rat Data Carbaryl was administered in the diet to Sprague-Dawley Crl:CD BR rats at 0, 250, 1500, and 7500 ppm (equating respectively to 0, 10.0, 60.2, and 349.5 mg/kg/day in males and 0, 12.6, 78.6, and 484.6 mg/kg/day in females) for at least 104 weeks (Hamada, 1993a). A subgroup of 10 rats/ sex/group was sacrificed after 52 weeks. Although survival was comparable across all groups, 7500 ppm exceeded the MTD primarily based on significantly lower body weight gains and food consumption. Compared to controls, body weight gains for the high-dose group at week 13 were reduced by 30 to 50% and by week 105, mean body weights were 35 and 45% lower for males and females, respectively. Mean body weights also tended to be lower at 1500 ppm and by week 105 were 6 and 12% below controls for males and females, respectively. Inhibition of RBC and brain ChE was observed at 1500 and 7500 ppm at both interim and final sacrifices. Increased incidences of microscopic findings related to treatment were observed only at 7500 ppm and consisted of transitional cell hyperplasia, papillomas, and carcinomas in the urinary bladder of males and females, pelvic epithelial hyperplasia in kidney of males with a single incidence of a kidney transitional cell carcinoma in one male, thyroid follicular cell hypertrophy in males and females with thyroid follicular adenomas in males only, and hepatocellular hypertrophy in males and females with hepatocellular adenomas in females only (Table 74.2). Increased incidences of urinary bladder transitional cell, epithelial hyperplasia in the renal pelvis, and hepatocellular hypertrophy were also observed at the 52-week interim sacrifice. Cell cycling assessments of the urinary bladder and thyroid gland from male rats and the liver from female rats of the control and high-dose groups at the 52-week sacrifice interval were performed using proliferating cellular nuclear antigen (PCNA) (Irisarri, 1996). Compared to control, a significant increase in the number of PCNA-positive urothelial cells was seen in the urinary bladder of the 7500 ppm males, while only slight increases were observed in the thyroid glands of males and the liver of females from the same treated group. The overall NOAEL for the study was 250 ppm.

74.3.6.2  Mouse Data

74.3.6  Carbaryl Chronic Toxicity and Carcinogenicity Studies Two-year studies with carbaryl administered via the diet have been conducted in rats and mice. In both of these studies, the highest dose tested exceeded the maximum tolerated dose (MTD), and thus, the findings described in the following sections for these high dietary exposures cannot be considered relevant for assessment of hazard or risk.

In a 2-year study in CD-1 mice, carbaryl was administered via dietary admixture at 0, 100, 1000, and 8000 ppm (equating to 0, 14.7, 146.0, and 1248.9 mg/kg/day in males and 0, 18.1, 180.9, and 1440.6 mg/kg/day in females) (Hamada, 1993b). Ten animals/sex/group were sacrificed after 52 weeks. Although survival was comparable across all groups, 8000 ppm exceeded the MTD based mainly on significantly lower body weight gains, clinical signs, and histopathological changes in bladder, kidneys,

Chapter | 74  Toxicological Profile of Carbaryl

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Table 74.2  Incidence of Microscopic Findings in the Rat Carcinogenicity Study with Carbaryl (All Animals on Study) Males Dietary levels in ppm Total number of animals

Females

0

250

1500

7500

0

250

1500

7500

70

70

70

70

70

70

70

70

Urinary bladder Transitional cell hyperplasia

8

7

11

51

6

4

4

56

Transitional cell papilloma

0

0

0

14

1

0

0

8

Squamous cell papilloma

0

0

0

2

0

0

0

0

Transitional cell carcinoma

0

0

0

10

0

0

0

5

Kidney Pelvic epithelial hyperplasia

13

10

13

29

22

39

29

21

Transitional cell carcinoma

0

0

0

1

0

0

0

0

Thyroid gland Follicular cell hypertrophy

2

1

1

9

3

4

2

33

Follicular cell adenoma

0

2

0

9

1

0

0

1

Follicular cell carcinoma

0

0

1

0

0

0

0

0

Liver Hepatocellular hypertrophy

0

1

2

38

7

6

10

34

Hepatocellular adenoma

1

1

1

1

1

0

3

7

Hepatocellular carcinoma

0

2

3

1

0

0

0

0

and spleen. Clinical signs of toxicity noted at 8000 ppm included hunched posture, thin and languid appearance, opaque eyes, urine stains, rough hair coat, soft feces, and low body temperature. Further, body weight gains were 33 and 19% lower than control for high-dose males and females, respectively, during the first 13 weeks of treatment and remained well below control gains for the entire study. Decreased RBC counts, hemoglobin concentration, and hematocrit were observed at 8000 ppm. RBC and brain ChE were decreased at 1000 and 8000 ppm at both the interim and final sacrifice intervals. Non-neoplastic findings observed after 52 and 104 weeks of treatment included intracytoplasmic protein-like droplets in urinary bladder superficial transitional epithelium at 1000 and 8000 ppm and increased incidence of lens cataracts and increased severity of extramedullary hematopoiesis and pigment in the spleen at 8000 ppm. Also, an increased incidence of chronic progressive nephropathy was observed at the 52-week sacrifice in males at 1000 ppm and in both sexes at 8000 ppm. Cell cycling assessments of liver and kidney from male and female mice of the control and high-dose groups at the 52-week sacrifice interval were performed using PCNA (Debruyne, 1998; Irisarri, 1996).

Only a slight increase in the number of PCNA-positive cells was observed in the kidney from treated males. Vascular neoplasms involving a variety of organs, with the liver and spleen being the most often affected, were observed in all treated male groups and in females at 8000 ppm (Table 74.3). Increased incidences of tubular cell neoplasms in the kidney and hepatocellular neoplasms were also observed at 8000 ppm in males and females, respectively. To better understand the carcinogenic potential of carbaryl, especially with regard to vascular tumors, a 6-month study was conducted with male C57BL/6 p53 heterozygous knockout mice. These mice carry only one wild-type p53 gene (p53 /) and are more susceptible to genotoxic carcinogens (Donehower, 1996; Tennant et al., 1995). In previous work, the genotoxic compound urethane, known to produce vascular tumors in lifetime studies in mice, also induced vascular tumors in p53 heterozygous knockout mice within 6 months of administration, and D-limonene, which has been identified as a nongenotoxic carcinogen in rat only (kidney tumors due to a specific male rat mechanism), was found to be negative (Carmichael et al., 2000). These results, along with the essentially zero level of vascular

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Table 74.3  Tumor Incidence in the Mouse Carcinogenicity Study with Carbaryl (All Animals on Study) Males

Females

Dietary levels in ppm

0

100

1000

8000

0

100

1000

8000

Total number of animals

80

80

80

80

80

80

80

80

Vascular Tumors Total hemangiomas

1

1

2

2

2

1

1

0

Total hemangiosarcomas

1

9

11

16

4

6

4

10

Total number of vascular tumors

2

10

13

18

6

7

5

10

Vascular tumor-bearing animals

2

7

10

10

4

4

4

9

Kidney Tubular cell adenoma

0

0

0

3

0

0

0

0

Tubular cell carcinoma

0

0

1

3

0

0

0

0

Liver Hepatocellular adenoma

12

6

11

10

1

0

1

7

Hepatocellular carcinoma

6

7

3

6

0

1

1

2

tumors in control p53 heterozygous knockout mice, support this model and study design as being suitable to investigate low-potency compounds suspected of being vascular carcinogens. Carbaryl was administered for 6 months to male C57BL/6 p53 heterozygous knockout mice via dietary admixture at 0, 10, 30, 100, 300, 1000, or 4000 ppm (equivalent to 0, 1.76, 5.21, 17.5, 51.6, 164.5, and 716.6 mg/kg/day, respectively) (Bigot-Lasserre et al., 2003; Chuzel, 2000b). The highest dietary level was selected based on a 28-day range-finding study in which p53 wild-type C57BL/6 male mice showed a 14% loss in body weight during the first week of treatment at 8000 ppm followed thereafter by little gain and a similar but less pronounced effect at 4000 ppm (Dange, 1998). No treatment-related mortalities or clinical signs were observed in the 6-month study. A slight decrease in body weight and food consumption was noted at 4000 ppm. The only nonproliferative change observed in the study was the accumulation of globular deposits in the umbrella cell layer of the urinary bladder at 100 ppm and higher with a dose-related increase in incidence and severity. This finding was not accompanied by any sign of local irritation or hypertrophy of the bladder epithelium. Histopathological examinations revealed no evidence of carbaryl-induced neoplasms of any type; in particular, no neoplastic or preneoplastic changes were noted in the vascular tissue of any of the organs examined. The only neoplasms observed were distributed randomly within groups and were recognized as occurring spontaneously in untreated mice of this strain. Thus, under the conditions of this study, the NOAEL of carbaryl was 4000 ppm (approximately 716 mg/kg/day) for

neoplastic changes. These results were not due to the lack of sensitivity of the model, since a previous study using urethane as positive control had demonstrated the sensitivity of the model for the induction of vascular tumors. Furthermore, a good correlation has been shown between the carcinogenic responses in the p53 knockout mouse model versus the conventional mouse bioassay, suggesting that most genotoxic chemicals could be detected in the p53 knockout mouse model (Storer et al., 2001). The negative results for carbaryl in p53 knockout mice along with the lack of evidence for binding to DNA or chromatid proteins show that a genotoxic mechanism is unlikely to be involved in the induction of vascular tumors in CD-1 mice by carbaryl.

74.3.7  Rat Metabolism and Toxicokinetics As the rat is the test species for the majority of toxicity studies conducted on a pesticide, understanding the metabolism and toxicokinetics of carbaryl in this species is important for the interpretation of the toxicological profile. In addition to general metabolism studies, work has been conducted to investigate carbaryl metabolism in older rats exposed to high levels of carbaryl via the diet as well as a limited investigation of metabolism in mice. The toxicokinetics of carbaryl have also been investigated following oral, intravenous (iv), and dermal dosing. The results from all of these studies are summarized below. Earlier work on the metabolism of carbaryl has been previously summarized and reviewed (Knaak, 1971).

Chapter | 74  Toxicological Profile of Carbaryl

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O O

OSO3H

HO Hydroxy des-methyl Carbaryl

OH

1-Napthol Sulfate

NH2

O O

OGlu

N H

1-Napthol

HO Hydroxy N-Hydroxymethyl Carbaryl O

1-Napthol Gluguronide O O

O N H

O

O

N H

Carbaryl

OH

OH

N H

OH

N-Hydroxymethyl Carbaryl

4-Hydroxy Carbaryl O O

O

O N H

O

OH OH

N H

HO OH

OH

5-Hydroxy Carbaryl

O

N H

3,4-Dihydro-3,4-dihydroxy Carbaryl

5,6-Dihydro-5,6-dihydroxy Carbaryl

OH

OH 1,5-Dihydroxy Naphthelene

Figure 74.1  Proposed metabolic pathway of carbaryl in the rat.

74.3.7.1  General Metabolism Data General metabolism work has been conducted in young Sprague-Dawley rats by Struble (1994). In these studies, a single dose of [14C]carbaryl was administered by oral gavage at 1 and 50 mg/kg or intravenously at 1 mg/kg. The metabolism of carbaryl following repeated exposure was also examined in rats administered a single oral dose of 1 mg/kg [14C]carbaryl following 14 consecutive days of pretreatment with non­ labeled carbaryl at 1 mg/kg. Urine and feces were collected for 168 h following dosing, at which time the animals were sacrificed for tissue collection. Metabolites were identified in urine and feces by either cochromatography with reference standards or isolation and identification by mass spectrometry. In all experiments, mass balance ranged from 96 to 104% of the administered dose, and no appreciable differences were observed in excretion rates or metabolism between the sexes.

The majority of the 1 mg/kg dose, administered either orally (single and repeated exposures) or intravenously, was excreted within 12 h of dosing, with 72 to 83% excreted in the urine and approximately 10% excreted in the feces. At 50 mg/kg, approximately 80% of the dose was excreted in urine and up to 12% in the feces, with nearly half the dose being excreted during the first 12 h postdosing and the greater part of the remainder excreted by 48 h postdosing. As the majority of the radiolabel was excreted within 24 h of dosing, very little radioactivity was detected in tissues collected at sacrifice 168 h after dosing. The main routes of metabolism were hydrolysis and oxidation, with both pathways generating hydroxylated metabolites that were then either excreted or conjugated prior to excretion. The metabolic pathway in the rat is shown in Figure 74.1. The major metabolites identified in excreta were 1-napthol

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and 5-hydroxycarbaryl. A small amount of parent compound was recovered in excreta. Several metabolites, 5,6dihydro-5,6-dihydroxy carbaryl, 3,4-dihydro-3,4-dihydroxy carbaryl, 4-hydroxycarbaryl, and 5-hydroxycarbaryl, were hypothesized as being formed via epoxide intermediates. Carbaryl metabolism has also been investigated to determine whether a dose-related shift in the metabolic profile occurred in older rats at the dietary levels used in the 2-year rat study that might account for tumor formation (Totis, 1997). In this study, groups of male CD (SpragueDawley-derived) rats received carbaryl via admixture in the diet at 0, 250, 1500, or 7500 ppm for 83 days followed by oral gavage of [14C]carbaryl at 2 mg/kg/day for 7 days. For purposes of comparison, a group of five male rats was administered a single oral dose of [14C]carbaryl at 50 mg/kg. All animals were 15 months old at study initiation compared to the 5- to 9-week-old rats used in the study by Struble (1994). As in the study in young rats by Struble (1994), the majority of the administered dose was excreted rapidly in the urine. For the dietary dosing regimens, urinary and fecal excretion totaled 96 to 103% of the administered dose, with most of the radioactivity being eliminated in the urine within 24 h after each administered radioactive dose. Mean levels for urinary elimination were 90, 90, and 66% of the dose administered, respectively, for the dietary groups of 250, 1500, and 7500 ppm. By comparison, following the single oral dose of 50 mg/kg, excretion in urine and feces accounted for 86 and 11% of the administered dose, respectively. Tissue levels were extremely low and represented 0.4% of the administered dose 168 h after the bolus dose of 50 mg/kg and 0.4 to 0.8% of the dose 3 days after the last administered radioactive dose in the groups initially receiving carbaryl via the diet. A shift in excretion was observed for two of the three major metabolites eliminated via the urine (Table 74.4). The glucuronide dihydro-dihydroxy 1-naphthyl-N-methylcarbamate metabolite, which is hypothesized to be formed by an epoxide intermediate, suggests that a higher proportion of reactive intermediates may have been present at the highest exposure level in the 2-year rat study. A limited metabolism study has also been conducted in CD-1 male mice (10 per group) fed diets containing 0, 10,

100, 1000, or 8000 ppm carbaryl for 14 days followed by a single gavage dose of 50 mg/kg [14C]carbaryl on the 15th day (Valles, 1999). Mean total radioactivity recovered, expressed as total administered dose, ranged from 88.7 to 101%, with greater amounts eliminated in the urine (55.8 to 68.9%) compared to the feces (12.2 to 18.6%). Less than 1% of the dose was found in the carcass 168 h after the radioactive dose. The four major components identified in the urine were the dihydro dihydroxy-naphthyl sulfate, the hydroxy-carbaryl glucuronide, -naphthyl sulfate, and -naphthyl -d glucuronide. The first two, possibly formed by epoxide intermediates, were increased in the mice given 8000 ppm in the diet, suggesting that at high doses of carbaryl, the metabolism, distribution, and excretion pattern may be altered, with a higher proportion of reactive intermediates being formed. Comparison with results from the rat suggests that there are some differences in metabolism, although a more complete study would be required to elucidate the metabolic pathway in mice.

74.3.7.2  Toxicokinetic Data To investigate toxicokinetics, [14C]carbaryl was administered to male Sprague-Dawley rats at two dose levels via oral gavage (1.08 or 8.45 mg/kg), dermal application to shaved skin for a 10-h exposure (17.25 or 102.95 mg/kg), or iv administration (0.8 or 9.2 mg/kg) (Krolski et al., 2003). Groups of four rats were sacrificed at various time points following dose administration, and total radio­active residue (TRR) levels were determined for whole blood, plasma, RBC, brain, liver, and fat. When sufficient residue was present in plasma or brain, metabolite profiles were determined and residues were characterized and/or identified. Rapid and complete uptake, metabolic degradation, and depletion of [14C]carbaryl were observed following oral and iv administration. Peak levels of radioactivity in brain, RBC, and liver were reached within approximately 15 min following oral and iv dosing, while radioactivity levels in fat peaked 30 to 60 min after dosing (Tables 74.5 and 74.6). For dermal exposure, only very low levels of radioactivity were detected in the matrices investigated. Peak levels in plasma were reached within 4 and 12 h for the low- and

Table 74.4  Levels of Major Urinary Metabolites Excreted by Male Rats (Approximately 18 Months Old) Urinary metabolitea

0 ppm

250 ppm

1500 ppm

7500 ppm

1-Naphthyl sulfate

24.14

27.19

22.93

11.68

Glucuronide of dihydro-dihydroxy 1-naphthyl-N-methylcarbamate

14.52

15.65

20.55

28.46

-naphthyl -d glucuronide

15.69

15.51

14.03

15.15

a

Expressed as percentage of dose administered.

Chapter | 74  Toxicological Profile of Carbaryl

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high-dose groups, respectively, and were approximately an order of magnitude lower than peak levels detected in plasma after oral and intravenous dosing. In all experiments, residues depleted rapidly. In plasma, parent compound was detected only within the first hour after iv administration at 9.2 mg/kg. Parent compound was also identified in brain, fat, and liver in rats receiving the oral and iv high-dose levels. In the brain, parent compound accounted for 70 to 90% of the total radioactive residue for up to 1 h after oral dosing and up to 2 h after iv dosing. By 2 h after oral and iv dosing, more than 90% of the initial level of carbaryl in the brain had been cleared. The relationship between the level

of carbaryl in the brain and ChE inhibition is shown in Figure 74.2. The hydrolysis product 1-napthol was identified in all tissues analyzed, while the sulfate conjugate of 1-napthol was found only in plasma. The oxidation product N-hydroxymethyl carbaryl was detected in brain only within 1 h after dosing. Ring-hydroxylated carbaryl was not observed in any tissue. The results of the oral and iv studies have been used to create a physiologically based pharmacokinetic and pharmacodynamic model (Nong et al., 2008). The model describes the tissue dosimetry of carbaryl and its metabolites (1-naphthol and “other hydroxylated metabolites”) as well as inhibition of ChE in RBC and brain.

Table 74.5  Average Total Radioactive Residue Levels (ppm) in Male Rats Following Oral Dosing Time (h)

Oral: 8.45 mg/kg

Oral: 1.08 mg/kg

Plasma

RBC

Brain

Liver

Fat

Plasma

RBC

Brain

0.25

7.19

2.56

1.97

20.95

3.58

1.44

0.44

0.13

0.5

7.71

2.59

1.15

13.45

3.38

1.19

0.32

0.06

1

7.32

2.24

0.62

  8.12

5.33

0.81

0.18

0.03

2

3.45

0.61

0.21

  2.57

1.27

0.54

0.10

0.03

4

1.59

0.36

0.11

  1.64

0.31

0.40

0.11

0.02

6

1.14

0.41

0.10

  1.90

0.24

0.19

0.07

0.01

12

0.53

0.15

0.06

  0.69

0.10

0.06

0.02

0.01

24

0.06

0.04

0.01

  0.14

0.02

0.01

0.01

0.00

Table reprinted from Nong et al. (2008) with permission of the publisher, Taylor & Francis Ltd.

Table 74.6  Average Total Radioactive Residue Levels (ppm) in Male Rats Following Intravenous Dosing Time (h)

Intravenous: 9.2 mg/kg

Intravenous: 0.8 mg/kg

Plasma

RBC

Brain

Liver

Fat

Plasma

RBC

Brain

0.083

11.74

10.18

13.22

24.67

12.07

2.13

1.06

0.74

0.167

12.36

  9.09

10.71

27.72

15.61

1.83

0.91

0.41

0.333

14.32

  7.34

  7.86

25.53

21.18

1.69

0.67

0.23

0.5

12.29

  5.51

  6.70

19.50

28.46

1.43

0.49

0.14

1

11.30

  3.23

  2.45

13.05

16.18

0.85

0.30

0.06

2

  7.78

  2.53

  1.09

  7.17

  8.48

0.58

0.15

0.03

4

  3.50

  1.04

  0.29

  2.70

  1.32

0.21

0.10

0.01

8

  1.50

  0.65

  0.15

  1.59

  0.21

0.11

0.05

0.01

Table reprinted from Nong et al. (2008) with permission of the publisher, Taylor & Francis Ltd.

Hayes’ Handbook of Pesticide Toxicology

1616

% ChE inhibition

100 80 60 40 20 0

0

2

4

6

8

10

12

14

Carbaryl level (ppm) Figure 74.2  Brain ChEI as a function of carbaryl level in the brain of male rats following IV dose of 9.2 mg/kg.

Conclusion Carbaryl belongs to the N-methyl carbamate chemical family. Carbaryl is moderately toxic by the oral route and exhibits a low level of toxicity following dermal or inhalation exposures. As with most carbamates, the cholinergic signs and symptoms associated with acute poisoning by carbaryl appear rapidly. In animal studies following single and repeated dosing, the carbamylation of ChE is reversible, and at oral exposures of 50 mg/kg or less, enzyme activity is similar to baseline values within 24 h after exposure. This finding is consistent with metabolism and toxicokinetic studies showing that the majority of a dose is rapidly metabolized and excreted within 24 h after either single or repeated administration. In neurotoxicity testing, the most sensitive endpoints were observations in the FOB, reduced motor activity, and ChEI in RBC and brain. The severity and frequency of clinical signs and ChEI were dose related and decreased with time. In genetic toxicity testing, some evidence of chromosomal aberrations has been observed in in vitro studies with carbaryl. However, all in vivo studies, including a 6-month feeding study in the p53 knockout mouse model, are clearly negative. In long-term feeding studies, neoplastic findings were observed in both rats and mice. In general, these findings were observed only at dose levels exceeding the MTD. However in male mice, the incidence of vascular tumors was elevated at all exposure levels. When tested in the p53 knockout mouse model, no evidence of carbaryl-induced neoplasms of any type, in particular, no neoplastic or preneoplastic changes, were noted in the vascular tissue of any of the organs examined. These results for carbaryl in p53 mice along with the lack of evidence for binding to DNA or chromatid proteins show that a genotoxic mechanism is unlikely to be involved in the induction of vascular tumors by carbaryl.

References Austin, E. (2002). 4-week repeated-dose dermal toxicity study with carbaryl technical in rats. Unpublished Report No. 6224-268. Covance Laboratories Inc.

Bigot-Lasserre, D., Chuzel, F., Debruyne, E. L. M., Bars, R., and Carmichael, N. G. (2003). Tumorigenic potential of carbaryl in the heterozygous p53 knockout mouse model. Food Chem. Toxicol. 14, 99–106. Brooks, W., and Broxup, B. (1995a). A time of peak effects study of a single orally administered dose of carbaryl, technical grade, in rats. Unpublished Report No. 97388. Bio-Research Laboratories Ltd. Brooks, W., and Broxup, B. (1995b). An acute benchmark-dose toxicity study of a single orally administered dose of carbaryl, technical grade, in rats. Unpublished Report No. 97387. Bio-Research Laboratories Ltd. Brooks, W., and Broxup, B. (1995c). An acute study of the potential effects of a single orally administered dose of carbaryl, technical grade, on behaviour and neuromorphology in rats. Unpublished Report No. 97389. Bio-Research Laboratories Ltd. Carmichael, N. G., Debruyne, E. L. M., and Bigot-Lasserre, D. (2000). The p53 heterozygous knockout mouse as a model for chemical carcinogenesis in vascular tissue. Environ. Health Perspect. 108, 61–65. Cifone, M. A. (1991). Mutagenicity test on carbaryl technical in the in vitro rat primary hepatocyte unscheduled DNA synthesis assay. Unpublished Report No. 10862-0-447. Hazleton Laboratories America, Inc. Chaung, W., Mi, L.-J., and Boorstein, R. J. (1997). The p53 status of Chinese hamster V79 cells frequently used for studies on DNA damage and DNA repair. Nucleic Acids Res. 25, 992–994. Cheng, T. (1994). Dermal absorption of 14C-carbaryl (80S) in male rats (preliminary and definitive phases). Unpublished Report No. HWI6224-207. Hazleton Wisconsin, Inc. Cheng, T. (1995). Dermal absorption of 14C-carbaryl (XLR Plus) in male rats (preliminary and definitive phases). Unpublished Report No. HWI6224-206. Hazleton Wisconsin, Inc. Chuzel, F. (2000a). Carbaryl 6-month carcinogenicity study in C57Bl/6 wild type mice by dietary administration. Unpublished Report No. SA 98217. Aventis CropScience. Chuzel, F. (2000b). Carbaryl 6-month carcinogenicity study in C57Bl/6 knock out mice by dietary administration. Unpublished Report No. SA 98155. Rhône-Poulenc Agro. Cranmer, M. F. (1986). Carbaryl: a toxicological review and risk analysis. Neurotoxicology 7, 247–332. Dange, M. (1998). Carbaryl preliminary 28-day toxicity study in the male TSG p53 wild type mouse by dietary administration. Unpublished Report No. SA 97499 and SA 97538. Rhône-Poulenc Agro. Debruyne, E. (1998). 52-Week toxicity study in the CD1 mouse target organs cell cycling assessment. Unpublished Report No. SA 97529. Rhône-Poulenc Agro. Donehower, L. A. (1996). The p53-deficient mouse: a model for basic and applied cancer studies. Semin. Cancer Biol. 7, 269–278.

Chapter | 74  Toxicological Profile of Carbaryl

Griffon, B. (2001a). Carbaryl: acute oral toxicity in rats. Unpublished Report No. 22010 TAR. Safepharm Laboratories Limited. Griffon, B. (2001b). Carbaryl: acute dermal toxicity in rats. Unpublished Report No. 22011 TAR. Safepharm Laboratories Limited. Griffon, B. (2001c). Carbaryl: acute dermal irritation in rabbits. Unpublished Report No. 22012 TAL. Safepharm Laboratories Limited. Griffon, B. (2001d). Carbaryl: acute eye irritation in rabbits. Unpublished Report No. 22013 TAL. Safepharm Laboratories Limited. Griffon, B. (2001e). Carbaryl: skin sensitization test in guinea pigs. Unpublished Report No. 22014 TSG. Safepharm Laboratories Limited. Gunasekara, A. S., Rubin, A. L., Goh, K. S., Spurlock, F. C., and Tjeerdema, R. S. (2008). Environmental fate and toxicology of carbaryl. Rev. Environ. Contam. Toxicol. 196, 95–121. Hamada, N. N. (1987). One-year oral toxicity study in Beagle dogs with carbaryl technical. Unpublished Report No. 400-715. Hazleton Laboratories America, Inc. Hamada, N. N. (1990a). Range-finding toxicity study in rats with carbaryl technical. Unpublished Report No. 656-137. Hazleton Laboratories America, Inc. Hamada, N. N. (1990b). Range-finding toxicity study in mice. Unpublished Report No. 656-136. Hazleton Laboratories America, Inc. Hamada, N. N. (1991). Subchronic toxicity study in dogs with carbaryl technical. Unpublished Report No. 656-152. Hazleton Laboratories America, Inc. Hamada, N. N. (1993a). Combined chronic toxicity and oncogenicity study with carbaryl technical in Sprague-Dawley rats. Unpublished Report No. 656-139. Hazleton Washington, Inc. Hamada, N. N. (1993b). Oncogenicity study with carbaryl technical in CD-1 mice. Unpublished Report No. 656-138. Hazleton Washington, Inc. Knaak, J. B. (1971). Biological and nonbiological modifications of carbamates. Bull. W. H. O. 44, 121–139. Krolski, M. E., Nguyen, T., Lopez, R., Ying, S.-L., and Roensch, W. (2003). Metabolism and pharmacokinetics of [14C] carbaryl in rats. Unpublished Report No. 201025. Bayer CropScience. Irisarri, E. (1996). 52-Week toxicity study in the rat and mouse target organs cell cycling assessment. Unpublished Report No. SA 95493. Rhône-Poulenc Agrochimie. Lawlor, T. E. (1989). Mutagenicity test on carbaryl (technical) in the Ames Salmonella/microsome reverse mutation assay. Unpublished Report No. 10862-0-401. Hazleton Laboratories America, Inc. Marshall, R. (1996). Carbaryl: induction of micronuclei in the bone marrow of treated mice. Unpublished Report No. 198/89-1052. Corning Hazleton. McEnaney, S. (1993). Study to evaluate the chromosome damaging potential of carbaryl technical by its effects on the bone marrow cells of treated rats. Unpublished Report No. RPF 5/RBM. Hazleton Microtest. Murli, H. (1989). Mutagenicity test on carbaryl technical in an in vitro cytogenetic assay measuring chromosomal aberration frequencies in Chinese hamster ovary (CHO) cells. Unpublished Report No. 108620-437. Hazleton Laboratories, Inc. Nong, A., Tan, Y., Krolski, M. E., Wang, J., Lunchick, C., Conolly, R. B., and Clewell, J. (2008). Bayesian calibration of a physiologically based pharmacokinetic/pharmacodynamic model of carbaryl cholinesterase inhibition. J. Toxicol. Env. Heal. A 71, 1363–1381. Onfelt, A., and Klasterska, I. (1983). Spindle disturbances in mammalian cells. II. Induction of viable aneuploid/polyploidy cells and multiple

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chromatid exchanges after treatment of V79 Chinese hamster cells with carbaryl. Modifying effects of glutathione and S9. Mutat. Res. 119, 319–330. Onfelt, A., and Klasterska, I. (1984). Sister-chromatid exchanges and thioguanine resistance in V79 Chinese hamster cells after treatment with the aneuploidy-inducing agent carbaryl  S9 mix. Mutat. Res. 125, 269–274. Padilla, S., Marshall, R. S., Hunter, D. L., and Lowit, A. (2007). Time course of cholinesterase inhibition in adult rats treated acutely with carbaryl, carbofuran, formetanate, methomyl, methiocarb, oxamyl or propoxur. Toxicol. Appl. Pharmacol. 219, 202–209. Repetto-Larsay, M. (1998). Carbaryl developmental toxicology study in the rat by gavage. Unpublished Report No. SA 98070. RhônePoulenc Agro. Robinson, K., and Broxup, B. (1996). A 13 week study of the potential effects of orally administered carbaryl, technical grade, on behavior, neurochemistry, and neuromorphology in rats. Unpublished Report No. 97390. Bio-Research Laboratories Ltd. Robinson, K., and Broxup, B. (1997). A developmental neurotoxicity study of orally administered carbaryl, technical grade, in rats. Unpublished Report No. 97391. Bio-Research Laboratories Ltd. Sagelsdorff, P. (1994). Investigation of the potential for protein-and DNAbinding of carbaryl. Unpublished Report No. CB93/52. Ciba-Geigy Limited. Storer, R. D., French, J. E., Haseman, J., Hajian, G., Legrand, E. K., Long, G. G., Mixson, L. A., Ochoa, R., Sagartz, J. E., and Soper, K. A. (2001). p53 / hemizygous knockout mouse: overview of available data. Toxicol. Pathol. 29(Suppl.), 30–50. Struble, C. B. (1994). Metabolism of 14C-carbaryl in rats (preliminary and definitive phases). Unpublished Report No. HWI 6224-184. Hazleton Wisconsin Inc. Tennant, R. W., French, J. E., and Spalding, J. W. (1995). Identifying chemical carcinogens and assessing potential risk in short-term p53deficient mice bioassays using transgenic mouse models. Environ. Health Perspect. 103, 942–950. Totis, M. (1997). Investigation of the metabolism of [14C]-carbaryl in the 15 month old male rat following chronic dietary administration. Unpublished Report No. SA95288. Rhône-Poulenc Agrochimie. Tyl, R. W., Marr, M. C., and Myers, C. B. (1999). Developmental toxicity evaluation of carbaryl administered by gavage to New Zealand White rabbits. Unpublished Report No. 65C-7297-200/100. Research Triangle Institute. Tyl, R. W., Myers, C. B., Marr, M. C. (2001) Two-generation reproductive toxicity evaluation of carbaryl (RPA007744) administered in the feed to CD (Sprague-Dawley) rats. Unpublished Report no. 65C-07407400. Research Triangle Institute. U.S. EPA. (2008). Amended Reregistration Eligibility Decision (RED) for Carbaryl. Valles, B. (1999). Investigation of the metabolism of [14C]-carbaryl following 14 days administration to the male CD-1 mouse. Unpublished Report No. SA97481. Rhône-Poulenc Agro. Wesson, C. M. (2001). Carbaryl: acute inhalation (nose only) toxicity study in the rat. Unpublished Report No. 282/618. Safepharm Laboratories Limited. Young, R. R. (1989). Mutagenicity test on carbaryl (technical) in the CHO/HGPRT forward mutation assay. Unpublished Report No. 10862-0-435. Hazleton Laboratories America, Inc.

Chapter 75

Aldicarb: Toxicity, Exposure and Risks to Humans Ann M. Blacker, Iain D. Kelly, Jennifer L. Lantz, Gary J. Mihlan, Russell L. Jones and Bruce M. Young Bayer CropScience, Research Triangle Park, North Carolina

75.1  Introduction Aldicarb is a carbamate insecticide with activity on a broad range of pests (insects, mites, nematodes). Aldicarb is the sole active ingredient in Temik brand Aldicarb Pesticide, sold and used almost exclusively as a 15% granule. Since its registration in 1970, aldicarb has been used in the United States on a variety of crops. Over the years, a large body of toxicity, environmental fate, residue, and monitoring data has been generated, which together with practical use experience has been used in developing stewardship programs and improving application methods that support the continued registration and safe use of this insecticide. This chapter reviews the current use and hazard and exposure profiles for the chemical and outlines their integration into a quantitative risk assessment. Aldicarb chemistry, biological characteristics, mammalian toxicity data, pharmacokinetics, and the determination of the point of departure (POD) for risk assessment are described. Advances in the methodology for incorporating the rapid reversibility kinetics of cholinesterase inhibition (ChEI) by aldicarb into a quantitative dietary risk assessment are discussed.

75.2  Description, use, and biological mode of action Technical aldicarb belongs to the N-methyl carbamate chemical family. The pure (technical) material is a white crystalline solid with a water solubility of approximately 6000 ppm at 25°C and is stable at room temperature. The empirical formula of aldicarb is C7H14N2O2S, with a molecular weight of 190.3. The structural formula is shown in Figure 75.1, and the Chemical Abstract Service (CAS) name and number are 2-methyl-2(methylthio)propanal O-[(methylamino) carbonyl]oxime and 116-06-30, respectively. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

Aldicarb is a soil-incorporated, systemic insecticide that is absorbed by the root system with subsequent translocation throughout the plant for control of chewing and sucking pests. Aldicarb is formulated solely as a dust-free granule and is not produced as a liquid. This type of formulation significantly reduces potential dermal and inhalation exposures, making the product much safer for occupational uses. In addition, application beneath the surface of the soil to a depth of up to several inches significantly reduces the negative impact on beneficial insects, fish, birds, and other wildlife because the product is not available for exposure.

75.3  Hazard characterization: derivation of the reference   dose (Rfd) 75.3.1  Summary of Aldicarb Toxicity Aldicarb has a very robust toxicity database including acute, chronic, developmental, reproductive, and neurotoxicity studies. Aldicarb has high acute toxicity, causing cholinergic symptoms due to the inhibition of acetylcholinesterase (AChE). These symptoms are dose-dependent, are rapidly reversible, and do not occur at expected human exposure levels. Aldicarb is neither genotoxic nor carcino­ genic. It does not cause developmental or reproductive effects in the absence of maternal toxicity. In rats, livestock, plants, and the environment, aldicarb is rapidly metabolized to aldicarb sulfoxide, then slowly converted to aldicarb sulfone (structural formula shown in Figure 75.1). Parent and the sulfoxide and sulfone metabolites may then be further metabolized to oximes and nitriles. Both the sulfoxide and sulfone are potent AChE inhibitors. Aldicarb and its sulfoxide and sulfone metabol­ ites are the residues of concern for risk assessment for all routes of exposure. 1619

Hayes’ Handbook of Pesticide Toxicology

1620

O

O N

N H

O

H3C

S

N

CH3

O

H3C

CH3 CH3

S

CH3 CH3

O N H

CH3

N

O H3C

S

O O Figure 75.1  Structural formula of aldicarb (left), aldicarb sulfoxide (middle), and aldicarb sulfone (right).

80 70 60

%ChEI

50 40 30 20 10 0 RBC

Brain

RBC

Male

Brain Female

Parent - 0.5 mg/kg

Sulfoxide - 0.5 mg/kg

Sulfone - 10 mg/kg

Figure 75.2  ChEI in male and female rats 1 h after a single oral bolus dose of aldicarb or its metabolites.

75.3.2  Acute Toxicity Aldicarb technical is highly toxic by the oral, dermal, and inhalation routes with an oral LD50 in rats of approximately 1 mg/kg (Myers et al., 1985; West and Carpenter, 1966), a dermal LD50 in rabbits of 5 mg/kg (Striegel and Carpenter, 1962), and an inhalation LC50 in rats of 0.0039 mg/l (Nachreiner et al., 1985). Aldicarb sulfoxide shows acute oral toxicity comparable to parent with an oral LD50 of 0.9 mg/kg in rats while aldicarb sulfone, with a rat oral LD50 value of 25 mg/kg, is less toxic than either aldicarb or aldicarb sulfoxide (Nycum and Carpenter, 1968). The level of ChEI in red blood cells (RBC) and brains from rats treated at approximately half the LD50 for aldicarb and its sulfoxide and sulfone metabolites is shown in Figure 75.2 (Brooks and Broxup, 1994).

75.3.3  Neurotoxicity Acute neurotoxicity studies have been conducted with aldicarb and its sulfoxide and sulfone metabolites. A single dose of aldicarb was administered by oral gavage to rats at

O

N H

CH3

CH3 CH3

0, 0.05, 0.1, and 0.5 mg/kg, and a functional observational battery (FOB) and motor activity were performed prior to dosing, 30 to 60 min after dosing and on study day 7 and 14 (Robinson et al., 1994). At 0.5 mg/kg, significant effects were observed in the FOB and motor activity shortly after dosing. FOB findings included increased incidences of salivation and lacrimation, tremors, and ataxic gait and decreases in locomotor activity, rearing, arousal, grip strength, body temperature, and motor activity. These findings were associated with approximately a 50% decrease in RBC and brain ChE activity measured 45 min after ­dosing (Table 75.1). At 0.1 mg/kg, no effects were observed in the FOB or motor activity, but RBC ChE was inhibited 31 to 47% and brain ChE was inhibited 10 to 16%. Aldicarb sulfoxide was tested in rats for acute neurotoxicity at 0, 0.05, 0.1, and 0.5 mg/kg while aldicarb sulfone was tested at 0, 1, 10, and 20 mg/kg (Brooks and Broxup, 1995a,b). Changes in the FOB and motor activity, similar to those seen for aldicarb, were observed at the highest dose tested for both metabolites as well as at the mid-dose for aldicarb sulfone. As with aldicarb, all of these findings were transient and no neuropathological changes were observed at study termination 14 days after dosing. In a subchronic neurotoxicity study, rats were administered aldicarb by oral gavage at 0, 0.05, 0.2, and 0.4 mg/kg/ day (Robinson et al., 1995). Neurobehavioral assessments performed during weeks 4, 8, and 13 showed tremors, salivation, and decreased motor activity at 0.2 and 0.4 mg/kg/ day and decreased grip strength at 0.4 mg/kg/day. In addition, a dose-related increase in pinpoint pupils was noted at all doses. Inhibition of RBC ChE was observed at all doses throughout the study, while brain ChE was inhibited throughout the study at 0.2 and 0.4 mg/kg/day. In a developmental neurotoxicity (DNT) study, female rats were dosed by oral gavage with aldicarb technical at 0, 0.05, 0.1, and 0.3 mg/kg/day from gestation day (GD) 6 through lactation day (LD) 10 (Weiler, 1995). Lower body weight gain was observed from GD6 through LD4 for dams at 0.3 mg/kg/day. High-dose pups showed lower body weight during the lactation and postweaning periods, while mid-dose male pups exhibited lower body weight only on LD7 and during postweaning. Neurobehavioral findings of hunched posture, tremors, excessive salivation and lacrimation, miosis, ataxia, and decreased rearing were observed in high-dose dams. No changes in either brain morphometry

Chapter | 75  Aldicarb: Toxicity, Exposure and Risks to Humans

Table 75.1  Aldicarb ChEI Dose Response in Rats 45 Min after a Single Oral Bolus Dose Dose (mg/kg)

Percent inhibition Males RBC

Brain

Females RBC

Brain

0.05

6

0

9

5

0.1

47

10

31

16

0.5

52

45

54

50

or neuropathology were noted in pups. Measurement of RBC ChE activity 2 h after dosing dams on GD7 and LD7 showed 27% inhibition for the 0.3 mg/kg group compared to controls. No inhibition of brain or RBC ChE was found in pups on LD4 or 10. The maternal no observed adverse effect level (NOAEL) was 0.05 mg/kg/day based on miosis at 0.1 mg/kg/day. The developmental NOAEL was 0.05 mg/kg/day based on postweaning body weight decrement, reduced hindlimb grip strength, and foot splay in the female offspring on postnatal day 35 (PND35).

75.3.4  Developmental and Reproductive Toxicity Aldicarb does not cause developmental or reproductive effects in studies in the absence of maternal (or parental) toxicity. In a developmental toxicity study, rats were administered doses of 0, 0.125, 0.25, and 0.5 mg/kg/day via oral gavage on GD6 through 15 inclusive (Tyl and Neeper-Bradley, 1988). Maternal toxicity was indicated by mortality, clinical signs, and reduced weight gain and food consumption at 0.5 mg/kg/day. Lower maternal weight gain and food consumption were also noted at 0.25 mg/kg/day. Gestational parameters were not affected. At 0.5 mg/kg/day, lower fetal weight was observed and fetal examinations revealed increased incidences for dilation of the lateral ventricles of the brain and reduced ossification in the sixth sternebra. Based on these findings, the NOAEL for fetal and maternal toxicity was 0.25 and 0.125 mg/kg/day, respectively. In a developmental toxicity study in rabbits, animals were administered doses of 0, 0.1, 0.25, and 0.5 mg/kg/day via oral gavage on GD7 through 27 inclusive (Leng et al., 1983). Maternal toxicity was seen at the upper two dose levels and consisted mainly of effects on body weight. No effects were observed on gestational parameters or fetal development. The maternal NOAEL was 0.1 mg/kg/day and the fetal NOAEL was 0.5 mg/kg/day. In a two-generation reproductive toxicity study, rats were exposed to aldicarb via dietary admixture at 0, 2,

1621

5, 10, or 20 ppm (ca. 0, 0.1, 0.25, 0.5, or 1.0 mg/kg/day) beginning 70 days prior to mating and then throughout mating, gestation, and lactation (Lemen, 1991). Lower body weights were seen for F0 and F1 adults and F1 and F2 pups at 20 ppm. Further, lower pup survival was noted in the high-dose group for both the F1 and F2 generations. Inhibition of RBC ChE occurred in F0 and F1 adults (both sexes) at 20 ppm and in F0 adult females at 10 ppm. The NOAEL for adult systemic toxicity was 5 ppm and the reproductive and offspring NOAEL was 10 ppm based on decreased pup weight and reduced viability. No effects on reproduction or on the offspring were seen in the absence of parental toxicity.

75.3.5  Sensitivity to Infants and Children The DNT, developmental toxicity, and reproduction studies provide no evidence of qualitative or quantitative susceptibility in the young (U.S. EPA, 2007). In these studies, designed to mimic the natural exposure pathways over the different stages of early life development, the fetus or neonate has an indirect exposure resulting from treatment of the dam either by gavage as in the DNT and developmental toxicity studies or via the diet as in the reproduction study. Under these exposure conditions, all NOAELs for the developing organism are higher or equivalent to NOAELs for maternal toxicity. However, administration of a single oral bolus dose in corn oil to rats of different ages showed a differential response for brain ChEI (Moser, 1999). The reported ED50 values for brain ChE in PND17, PND27, and adult (approximately PND70) rats were 0.12, 0.19, and 0.29 mg/kg, respectively, for males and 0.15, 0.21, and ������������������������������������������������� 0.28 mg/kg��������������������������������� , respectively, for females. For neurobehavioral endpoints, young rats showed fewer clinical signs of ChEI compared with adults, and decreased motor activity was noted only in the adult animals. Administration of a bolus dose directly to pups at levels that produce a high degree of ChEI is not predictive of the effects associated with lower levels that might occur with relevant dietary exposures. Based on in vitro data, the differential response in brain ChEI following a single oral bolus dose does not appear to be due to dynamic differences in AChE sensitivity, as IC50 values for aldicarb were similar with PND4 and adult brains (Mortensen et al., 1998). Rather, the generally lower capacity of metabolic enzymes in neonatal animals may result in a greater internal exposure for the same administered dose.

75.3.6  Genotoxicity Studies covering gene mutations, chromosomal aberrations, and unscheduled DNA synthesis were all negative for aldicarb (Cimino et al., 1984; Godek et al., 1980a, 1984a; Ivett, 1990; Stankowski et al., 1985a). A limited battery of

1622

genotoxicity studies on aldicarb sulfoxide and sulfone are also negative (Godek et al., 1980b,c, 1984b; SanSebastian et al., 1984; Stankowski et al., 1985b).

75.3.7  Chronic Toxicity and Carcinogenicity Aldicarb produced no carcinogenic effects when administered to rats and mice in lifetime experiments. ChEI was the most sensitive endpoint in chronic studies in rats and dogs. Results of chronic toxicity and carcinogenicity testing are summarized below. In a National Cancer Institute study (NCI, 1979), rats and mice were fed aldicarb in the diet at 0, 2, and 6 ppm (equivalent to ca. 0, 0.1, and 0.3 mg/kg/day in rats and 0, 0.29, and 0.86 mg/kg/day in mice) for 103 weeks. No effects on mortality or body weight attributable to aldicarb were observed in either rats or mice. It was concluded that under the conditions of the assay, aldicarb was not carcinogenic. In a 2-year study, rats were fed aldicarb at levels of 0, 1, 10, and 30 ppm in the diet (equivalent to ca. 0.05, 0.5, and 1.5 mg/kg body weight/day) (Trutter, 1993). There were no compound-related effects on survival. The high dose of 1.5 mg/kg/day was greater than the oral LD50 and was tolerated every day over the course of the study. Administration via the diet would result in the total daily dose being ingested in fractionated amounts throughout the day, allowing for reversibility of ChEI between consumption periods. Significant decreases in RBC ChE were observed for the mid- and high-dose groups throughout the study. The principal clinical effect observed was limited use of the tail in high-dose males and females. Body weights and body weight gains were reduced in high-dose males and females. Also, abnormalities of the iris were reported for this dose group. No evidence of direct organ toxicity or carcinogenicity was observed. The NOAEL was 0.05 mg/kg/day based on RBC ChEI. In another 2-year study, rats were fed aldicarb at dose levels of 0 and 0.3 mg/kg (body weight)/day. In addition, other groups were fed aldicarb sulfoxide at dose levels of 0, 0.3, and 0.6 mg/kg/day, aldicarb sulfone at dose levels of 0, 0.6, and 2.4 mg/kg/day, or a mixture of aldicarb sulfoxide and aldicarb sulfone at doses of 0, 0.6, and 1.2 mg/kg/day (Weil and Carpenter, 1972). Neither aldicarb nor either of its major metabolites was found to be carcinogenic. There were slight increases in mortality and slight depressions in growth at certain stages for some of the test materials. ChE activity was measured at 6, 12, and 24 months during the study. However, as RBC and brain ChE activity were examined 24 h after animals were removed from test diets, no ChEI was noted for these matrices. In an 18-month study, mice were fed aldicarb at dose levels of 0, 0.1, 0.3, and 0.7 mg/kg (body weight)/day (Weil and Carpenter, 1974). There was no effect on mortality or growth. Inclusion of aldicarb in the diet did not result in an increased incidence of carcinogenic response.

Hayes’ Handbook of Pesticide Toxicology

In a 1-year study, dogs were administered aldicarb via the diet at 0, 1, 2, 5, and 10 ppm (equivalent to ca. 0, 0.027, 0.055, 0.13, and 0.24 mg/kg/day) (Hamada, 1988). ChE activity was measured from blood samples approximately 2 h after the 2-h feeding period during weeks 5, 13, 26, and 52. There were no observable effects other than ChEI. Statistically significant differences were observed for RBC ChE activity at the two highest doses and for brain ChE activity at the highest dose only. Thus, the NOAEL for RBC and brain ChEI was 0.055 and 0.13 mg/kg/day, respectively.

75.3.8  Human Volunteer Studies In a series of studies reported in 1973, groups of four adult male volunteers were administered aldicarb orally in aqueous solution at dose levels of 0.025, 0.05, and 0.1 mg/kg body weight. Clinical signs were recorded and whole blood ChE activity was measured up to 6 h after administration of the sample. Total urine voided was collected and aldicarb excretion patterns for the initial 8 h after dosing were evaluated. In addition, spot samples were taken at 12 and 24 h. In another investigation, two additional subjects were administered aldicarb in water at dose levels of 0.05 and 0.26 mg/kg. Dose levels of 0.1 and 0.26 mg/kg are considered to be high doses, and clinical signs, typical of anticholinesterase agents, were observed at these doses within 1 h of aldicarb administration. ChEI at these very high dose levels was observed in all volunteers within 1–2 h after treatment. Recovery was evident within the first 6 h after treatment. No evidence of toxicity was observed at the 0.05 mg/kg. Urine analysis showed that approximately 10% of the administered dose was excreted as carbamates within the first 8-h interval. In 1992, a double-blind, placebo-controlled study in human volunteers was conducted and performed according to existing and globally accepted ethical guidelines (Wyld et al., 1992). This study was reviewed by EPA’s Human Studies Review Board (U.S. EPA, 2006), which concluded that the ChE data were reliable for use in the aldicarb single-chemical, aggregate risk assessment from both a science and an ethical standpoint. In this study, aldicarb was administered as a single oral dose to healthy male and female subjects. The doses administered were as follows: placebo (22 subjects: 16 males and 6 females); 0.01 mg/kg (8 males); 0.025 mg/kg (8 males and 4 females); 0.05 mg/kg (8 males and 4 females); and 0.075 mg/kg (4 males). Volunteers were screened before entry for general medical history by examination and laboratory tests including hematology, clinical chemistry, and urinalysis. Clinical measurements were made at intervals before and after dosing. These included vital signs, pulmonary function tests, electrocardiographs (ECGs), and documentation of any signs or symptoms. Samples were taken for urinalysis, clinical chemistry (including RBC ChE activity), and hematology before and after dosing. There were no clinically significant changes

Chapter | 75  Aldicarb: Toxicity, Exposure and Risks to Humans

in vital signs, pupil size, pulmonary function, ECGs, salivation, clinical signs, clinical chemistry (apart from ChE), hematology, or urinalysis in the study. RBC ChE activity was significantly depressed 1 h after dosing at 0.05 and 0.075 mg/kg. ChEI and time to recovery were dose related with ChE activity returning to baseline by 8 h in all subjects. No biologically significant depression of RBC ChE activity was seen in subjects treated with 0.01 or 0.025 mg/kg. A single volunteer (0.075-mg/kg group, actual dose 0.06 mg/kg) reported generalized sweating. Thus, the NOAEL for clin­ ical signs was reported as 0.05 mg/kg and the NOAEL based on RBC ChEI was 0.025 mg/kg.

75.3.9  Pharmacokinetics of Cholinesterase Inhibition The pharmacokinetics of ChEI has been studied in both the rat and human. The results from both species have been subjected to a rigorous statistical assessment and, taken together, provide a detailed understanding of the time course of ChEI that can be used to further characterize the risk assessment for this molecule. In adult male rats, a two-phased experiment was conducted in which RBC ChE was monitored following single and double administrations of aldicarb by gavage. The phase I experiment was a single oral dose study with blood samples taken via jugular cannula prior to exposure and at 11 postadministration time points (from 10 to 240 min)

1623

for determination of ChE in RBC. The phase II experiment was similar, but included a second administration approximately 4.5 h after the first. The study was conducted at dose levels of 0, 0.05, and 0.10 mg/kg in both phase I and II. In phase I, there was rapid inhibition of ChE for both dose groups, followed by a recovery to baseline ChE over the next 1 to 3 h. The phase II data were similar with inhib­ ition and recovery following each dose administration. The statistical assessment of the data is described in detail in Clayton et al. (2003). A nonlinear mixed model approach, based on simple first-order kinetics, coupled with Monte Carlo simulation offered a method to assess the pharmacokinetics of ChE activity following exposure to aldicarb and provided an appropriate description of both the inhibition of ChE and the subsequent recovery phase. Several properties of the model are important in characterizing aldicarb effects within the range of doses used in the experiment (Figure 75.3). The time to maximum inhibition is not dependent on dose. For a single dosing, the reduction in ChE at any given time is proportional to the dose. Virtually all animals had recovered to near preexposure levels by the time of the second exposure. The estimated ChE rapidly declines and then exhibits a steady rate of recovery to the baseline level. The model predicts a second decline in ChE that starts from the level to which activity has recovered after the first administration and then ­continues the recovery process toward the baseline level with each aldicarb exposure event acting independently.

900 850

mU/mL of Packed RBC

800 750 700 650 600 550 500

Baseline 0.05 mg/kg 0.10 mg/kg 90% of Baseline

450 400 0.0

1.0

2.0

3.0

4.0

5.0

6.0

7.0

8.0

9.0

10.0

Time After Initial Administration (Hrs) Figure 75.3  Estimated time patterns of typical ChE activity in male rats administered an oral bolus dose of aldicarb at time 0 followed by a second oral bolus dose 4.5 h later (reprinted with permission from the Journal of Agricultural, Biological & Environmental Statistics. Copyright 2003 by the American Statistical Association. All rights reserved).

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The pharmacokinetics data in humans are more limited but consistent with the findings in rats. Data from the recovery phase of RBC ChEI in humans from the study of Wyld et al. (1992) were analyzed as described by Williams et al. (2008). The human data are from a single exposure with the first postexposure ChE measurement taken at 1 h. As seen in the rat study described previously, the peak inhibition occurs prior to 1 h postexposure. Thus, the portions of model relating to the inhibition pattern, the aggregate of absorption, binding, and carbamylation, cannot be estimated for the human study. The model, however, can be modified to estimate the inhibition at 1 h postexposure, followed by the recovery pattern for decarbamylation. As with the rat data, a key property of the model is that the recovery rate is not affected by administered dose; that is, the rate of recovery for all doses is the same. A parsimonious nonlinear mixed effects model, based on simple firstorder kinetics, with only three fixed effect parameters, was found to represent the dose–response relationship of the inhibition at 1 h postexposure and recovery of ChE activity thereafter. One of these parameters (3) represents the first-order elimination coefficient describing the rate of recovery and has been used as described in Section 75.4.5 to refine the dietary risk assessment for aldicarb.

75.3.10  Critical Effect and Point   of Departure Based on the extensive toxicology database for aldicarb, ChEI is identified as the most sensitive endpoint consistently across all species. This endpoint is the key event in carbamate toxicity and has been used as the basis for the point of departure (POD) throughout most of the risk assessment history of aldicarb. In general, the RBC activity is a more sensitive endpoint than brain for ChEI in response to aldicarb and is considered an appropriate surrogate of potential effects on peripheral nervous system AChE activity and of potential effects on the central nervous system when brain data are lacking (U.S. EPA, 2000). The rodent and human responses to RBC ChEI for aldicarb are similar. Further, the human RBC ChEI observed in both sexes at the common dose levels of 0.025 and 0.05 mg/kg suggests no differences between sexes for this response in humans. Therefore, RBC ChE activity from the human study is considered the most appropriate endpoint for derivation of the POD. As NOAELs can be influenced by dose selection and do not necessarily reflect the relationship between dose and response for a given chemical, dose–response modeling is an alternative approach for derivation of a POD (Crump, 1984). The 1998 Joint FAO/WHO Meeting (WHO, 1999) on Pesticide Residues proposed 20% RBC ChEI as an appropriate level for determining an endpoint protective of human health. However, EPA’s Office of Pesticide Programs uses a value of 10% RBC ChEI, as

Hayes’ Handbook of Pesticide Toxicology

they consider this level to be protective for other toxicities, such as clinical signs and/or behavioral endpoints, and it is generally at or near the limit of sensitivity for discerning a statistically significant decrease in ChE activity across the blood and brain compartments. The EPA estimated the benchmark dose at which 10% ChEI is observed (BMD10) and the lower 95% confidence interval (BMDL10) by fitting the ChE data to an exponential dose–response model using generalized nonlinear least squares. The BMD10 and BMDL10 estimates for the 1992 human RBC ChE data for aldicarb are 0.02 and 0.013 mg/kg, respectively (U.S. EPA, 2007).

75.4  Exposure characterization This section describes the quantitative data available to assess the nonoccupational routes of exposure to aldicarb, which can then be integrated with the hazard information described previously to estimate the potential risk to the general population and selected subpopulations. Aldicarb has no residential uses, and thus the routes of exposure are dietary consumption of residues in food or drinking water. As residue levels are generally determined in raw agricultural commodities (RACs) and not in foods as eaten, food consumption is transformed into RAC consumption using a series of translation recipes so that RAC residue values can be used directly in calculating dietary exposure. Each of the data sources described in the following sections provides quantification of the residues of concern, namely, parent aldicarb and the metabolites aldicarb sulfoxide and aldicarb sulfone. For simplification of the dietary risk assessment, residues are generally expressed as parent equivalents, a conservative assumption, as aldicarb and aldicarb sulfoxide are of approximately equal toxicity while the aldicarb sulfone is less toxic. In the discussion of risk, the term aldicarb will be used to represent the summation of parent and the two relevant carbamate metabolites.

75.4.1  Food Consumption Data The most comprehensive information on food consumption data for the U.S. population and subpopulations is the USDA’s Continuing Survey of Food Intakes by Individuals (CSFII) for the period 1994–1996 and the 1998 Supplemental Survey (USDA, 2000). The CSFII is a nationwide survey conducted by USDA’s Agricultural Research Service that provides the food intakes of 20,607 individuals of all ages on two nonconsecutive days. Data collection in 1994 to 1996 included individuals of all ages. The supplemental survey conducted in 1998 included children from birth to 9 years of age. The survey indicates the time of day food and/or a meal was consumed, which, given the rapid reversibility of aldicarb ChEI described in

Chapter | 75  Aldicarb: Toxicity, Exposure and Risks to Humans

Sections 75.3.9 and 75.4.5, in principle allows the hazard and exposure to be characterized by each meal or eating event. As stated previously, this survey provides information on food as consumed and not on food commodities. The Food Commodity Intake Database (FCID) provides the companion data (translation recipes) that allow the CSFII food consumption data to be expressed as RACs or processed commodities (U.S. EPA and USDA, 2000). For example, noodles will be translated to wheat flour and whole eggs, for which pesticide residue data is available.

75.4.2  Summary of Residues in Food Residue levels in foods as finally eaten by consumers can vary dramatically from those in or on treated crops at harvest. Dietary exposure incorporates three types of data to account for this variation: measured residue levels in the crop, commercial or home processing factors, and the percent of a crop that is treated in the United States. Feeding studies in ruminants and poultry, in which aldicarb is fed via diet to the animals and the tissues are then analyzed, indicate that aldicarb residues are not expected to occur in meat, milk, poultry, or eggs based on the registered uses of the product. Aldicarb residue data is available from a range of sources. Residue data for RACs and some processed commodities are generated from controlled trials conducted under label conditions that result in the most conservative measure of potential residues in the analysis. These include dry beans, pecans, peanuts, sugar beets, coffee, cotton, sorghum, soybeans, and sugar cane for aldicarb registered crops. The average field trial residues for these crops range from 0.007 mg/kg for sugar cane to 0.145 mg/kg for pecans. USDA Pesticide Data Program (PDP) monitors residues on major foods consumed by infants and children. The samples are collected from nationally representative large distribution centers and provide data on residues closer to consumption than field trial data. Sweet potato data in this analysis were from the PDP program. Out of 926 samples there were five detects ranging from 0.02 to 0.16 mg/kg. USDA PDP also conducted a special study that measured aldicarb residues in composite and individual potato samples. There were 20 out of 342 composite samples collected with detectable residues (range 0.012 to 0.172 mg/kg). The individual tuber data from 16 of the 20 composites with detectable residues had 96 out of 160 samples with detects ranging from 0.012 to 0.372 mg/kg. Finally, market basket survey data were used for citrus commodities and represent the residue data closest to consumption, as samples are collected from nationally ­representative grocery stores. There were 399 individual orange samples collected with 16 showing detects ranging from 0.001 to 0.0025 mg/kg. The PDP monitoring and market basket data measure residue levels after samples have been treated as they typically would in the home before consuming (e.g., potatoes were washed, oranges were peeled).

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Typical commercial processing factors from GLP-conducted studies are available for many of the crops such as coffee (roasted), cottonseed (oil), and orange (juice). In almost all cases reduction occurred from processing, the exception being dried citrus peel. For potato, a key driver in the acute risk assessment, the typical commercial factors of flakes, chips, and dried were used. In addition, studies for home factors of baking, frying, and boiling were conducted. All potato factors showed residue reduction. Finally, not all of an aldicarb registered crop will be treated with the pesticide. This dilutes the potential contribution to dietary exposure. Percent of the crop treated information is available from government and proprietary sources and is used to adjust the residue distributions.

75.4.3  Water Consumption Data Dietary consumption of water occurs from various sources. Water found naturally in foods is clearly accounted for in the food consumption databases. Additionally, the CSFII database contains water consumption data for indirect water, that is, water added to foods and beverages during final preparation at home or by cafeterias and restaurants, and for water consumed directly. The FCID database incorporates the recipe file to specify foods assumed to contain indirect water. For both direct and indirect water, the source of water was identified as tap (community water), bottled, other source (wells), or missing. The CSFII collected only the total daily consumption of direct water and did not record time of day of the individual water ­consumption events. A nationally representative water consumption ­survey has been conducted to address how often, when, and how much water is consumed at specific times during the day (Barraj et al., 2008). This data can be incorporated into the risk assessment to give the time of day information for water consumption and event-based analysis as with the CSFII foods data.

75.4.4  Summary of Residues in Drinking Water There is a potential for aldicarb to be found in drinking water from both surface water and groundwater sources. Aldicarb residues in surface water have been measured in a number of monitoring programs. These include the National Water Quality Assessment (NAWQA) program conducted by the U.S. Geological Survey (USGS), monitoring of aldicarb in intermittent streams in Tennessee by the USGS (Williams and Harris, 1996), surface water monitoring in California, and a targeted drinking water monitoring program (Jones, 2005). Since applications of aldicarb granules are incorporated into the soil, the results of surface monitoring studies show that residues in surface water are relatively rare and at low levels except in those

1626

cases where water can move downward into soil and then laterally via interflow on a more impermeable soil layer to small ditches or streams (one example is bedded citrus in southern Florida). Residues in larger streams in these areas where interflow is important are much smaller. The most comprehensive surface water study is a 3-year study monitoring aldicarb residues in raw and finished water in 28 different community water systems using surface water in aldicarb use areas in the United States. (Jones, 2005). The selection of community water systems was based largely on aldicarb use in the watershed, with the objective of selecting systems where the use intensity (amount of aldicarb applied/area of watershed) in the watershed was highest for that particular use pattern. The results show that overall residues were highest in southern Florida, associated with the use of aldicarb on citrus, where three of the six community water systems had low levels of aldicarb carbamate residues during significant portions of the year (Jones, 2005). The maximum concentration observed in finished water was 0.16 g/l and the maximum time-weighted annual average was 0.068 g/l. Residues were also present at lower concentrations and for shorter periods of time in other areas of the Southeast. Aldicarb carbamate residues in other portions of the country were insignificant. Aldicarb residues in groundwater have been measured in a number of monitoring programs. These include numerous state monitoring programs, national monitoring programs conducted by the EPA as well as the NAWQA program conducted by the USGS, and a variety of registrant monitoring and research studies. The work from 1979 through the mid-1980s shows that aldicarb residues have the potential to move into groundwater under certain conditions, but that residues in drinking water can be prevented by a number of best management practices including use restrictions in highly vulnerable areas, optimization of timing and rates of application, and well setbacks (Jones and Estes, 1995). A recent, comprehensive study of residues in drinking water from potable wells (Jones and Allen, 2007) has been used in dietary assessments. This study consisted of collecting potable well samples from nine major use areas located in five regions of the United States. The study included all major use areas with the exception of Florida. Samples were analyzed from 1673 drinking water wells, which were within 1000 feet of fields treated at least once with TEMIK brand 15G Aldicarb during 2002–2005. Samples from 1513 of the 1673 potable wells had no detectable aldicarb carbamate residues. Only 10 wells had total aldicarb carbamate residues above 1 g/l, the maximum being 2.9 g/l. All residues were below the U.S. EPA Health Advisory Limit of 10 g/l (U.S. EPA, 2004a).

75.4.5  Modeling of Exposure Scenarios Risk assessment integrates the components of hazard and exposure to make risk estimates for the exposure scenarios of interest. For aldicarb, RBC ChEI is a sensitive indicator of

Hayes’ Handbook of Pesticide Toxicology

hazard that is protective of other toxicities (U.S. EPA, 2007). Inhibition by aldicarb is rapidly reversible (Section 75.3.9) with an estimated half-life (t1/2) of 1.7 h based on the firstorder elimination coefficient 3  0.9 h1 (Williams et al., 2008) using ChEI from a human volunteer study (Wyld et al., 1992). As the magnitude of inhibition caused by an exposure event is independent of prior exposure, the exposure scen­ arios of interest are the individual consumption events (e.g., meal, snack, water ingestion) occurring within a day. While within-day consumption data is available as described in Sections 75.4.1 and 75.4.3 for food and water, respectively, current population-based, dietary risk assessment models are not fully configured to conduct within-day assessments that aggregate food and water exposure. There are three stochastic models (DEEM, LifeLine, CARES) commonly used to conduct dietary risk assessments that are well validated and publicly available (Exponent, Inc., The LifeLine Group, Inc., and International Life Sciences Institute). The three models give comparable results for dietary exposure of pesticides (U.S. EPA, 2004b). Stochastic methods combine the magnitude of each variable (e.g., food and water residues, consumption amounts) with its probability of occurrence to develop a frequency distribution of exposure. Factors that affect these variables, such as food-processing factors that affect residue levels and percent crop treated information that affects the probability of occurrence, can generally be incorporated directly into the model. All three available models are configured to conduct acute risk assessments by determining exposure distributions based on total daily consumption values, that is, the consumption of all identical food items (or water) summed for each day. The CARES model has released a Dietary Minute Module considering exposure based on time intervals of less than 1 day, although the within-day assessment concepts described below could be applied to any model, for any compound exhibiting similar kinetics. The Dietary Minute Module incorporates the time of day of food consumption from the CSFII and, using the FCID translation recipes, gives a consumption profile that shows by the minute the RACs and processed commodities (events or meals) consumed for the individuals in a population within each day (Kelly et al., 2006). While a full description of the issues associated with conducting a within-day risk assessment for aldicarb is beyond the scope of this chapter, the principles that have been incorporated into the Dietary Minute Module are worth describing. For rapidly reversible carbamates such as aldicarb, the rate of recovery following peak inhibition is essentially independent of dose and follows a first-order decline determined by the rate of decarbamylation of ChE and the rate of elimination of the carbamate from the body (Williams et al., 2008). Thus the biologically effective dose (or residual exposure) is the mirror image of ChE inhib­ ition during enzyme recovery described in Figure 75.3. Based on the exposure at each event, the biologically equivalent exposure can be calculated if the first-order

Chapter | 75  Aldicarb: Toxicity, Exposure and Risks to Humans

decline rate and the time from exposures are known [Eq. (1)]. The biological exposure at time t is the sum of all residual exposures, that is, the sum of each exposure adjusted by the ChE recovery since the exposure occurred. Di  Ei  ( Di1  exp[ exp(3 ) /60  lag ]) (1)



where Di is the biologically equivalent exposure at event i (mg/kg), Ei is the exposure at event i (mg/kg), 3 is the first-order recovery rate of ChE activity in h–1, and lag is the time since the last eating event (ti  ti1) in min. Thus if a series of exposure events occur as described in Table 75.2, the biological equivalent exposures can be calculated. The exposure profile for such a series of withinday exposures, compared to the daily exposure profile is illustrated in Figure 75.4. While the Dietary Minute Module allows the minuteby-minute profile of aldicarb exposure for an individual to be developed, and the information and matching algorithms are available to do the same for water, the development of a statistically weighted, population-based within-day assessment as mentioned previously is beyond the scope of this chapter. The results of the risk assessment described in Section 75.5 will focus on the daily assessment.

75.5  Risk characterization Risk characterization incorporates an integrative analysis that brings together the assessments of hazard, dose– response, and exposure to make risk estimates for the

1627

exposure scenarios of interest. Any percentile of exposure from the probabilistic distribution could be used and compared to an appropriate hazard endpoint. In the following discussion, the percentiles of exposure shown are the 95th, 99th, and 99.9th percentile. The endpoint of comparison is the acute reference dose (aRfD). For aldicarb, the aRfD is based on the BMDL10 of 0.013 mg/kg bw from the human study discussed previously with an intraspecies uncertainty factor of 10. The interspecies uncertainty factor is 1, as the point of departure is based on a human study. The typ­ ical daily dietary acute results using the CARES model are described below.

75.5.1  Acute Dietary (Food-Only) Risk Assessment Aldicarb is currently registered for use on citrus, coffee, cotton, dry beans, grain sorghum, peanuts, pecans, potatoes, soybeans, sugar beets, sugarcane, sweet potatoes, seed alfalfa, and tobacco. Each of these items, along with water, may be consumed as part of the diet by a member of a defined population or subpopulation and may contain residues of aldicarb. This analysis used the data and assumptions described in Sections 75.4.1 and 75.4.2 for the food-only assessment. The estimated risk from a 24-h acute dietary exposure analysis (99.9th percentile) is 37% of the aRfD for children 1–2 years, the most highly exposed subpopulation (Table 75.3). This is well below the threshold of concern of 100% of the RfD for most regulatory bodies.

Table 75.2  Hypothetical Within-Day Food and Water Exposure Events Source of exposure

Minute of daya

Time between last eventb

Event exposure (mg/kg/day)

Biologically equivalent dose (mg/kg/day)c

Water

390

390

2.85E-04

2.85E-04

Food

420

30

6.73E-04

9.05E-04

Food

615

195

3.06E-03

3.30E-03

Water

690

75

4.27E-04

2.41E-03

Water

720

30

7.62E-04

2.73E-03

Water

780

60

6.29E-04

2.45E-03

Water

1020

240

3.03E-04

7.84E-04

Water

1050

30

5.69E-04

1.21E-03

Food

1080

30

3.06E-03

4.05E-03

Water

1170

90

2.85E-04

2.48E-03

Food

1200

30

6.73E-04

2.70E-03

a

Minute of day (i.e., 720 minute  6:00 AM). Time between events (lag)  ti  t1i where i is a given event and t is time in minute of day. c Biologically equivalent dose (D)  Di  Ei  (Di1  exp[exp(3 )/60  lag]), where 3  0.9h1; E  event exposure (mg/kg/day). b

Hayes’ Handbook of Pesticide Toxicology

1628

Total Exposure By Day Biological Equivalent Exposure within day Biological Equivalent Exposure at each event Food Exposure Events

Exposure

Water Exposure Events

Time Figure 75.4  Within-day biological equivalent exposure compared to total exposure for the day.

Table 75.3  Daily Dietary Food Risk Assessment Population

Percentiles (weighted per capita) 95th

99th

99.9th

Exposure (mg/kg bw)

Percent aRfDa

Exposure (mg/kg bw)

Percent aRfDa

Exposure (mg/kg bw)

Percent aRfDa

Infants (1 year)

8.97E-06

1

6.16E-05

5

2.37E-04

18

Children (1–2 years)

2.23E-05

2

1.00E-04

8

4.78E-04

37

Adults (20–49 years)

6.50E-06

1

3.20E-05

2

1.71E-04

13

a

Percent aRfD calculated based on RfD  0.0013 mg/kg (BMDL10  0.013 mg/kg; UF  10).

75.5.2  Acute Dietary (Water-Only) Risk Assessment

75.5.3  Aggregate (Food-Plus-Water) Risk Assessment

Several surface water and groundwater dietary analyses were conducted using the available monitoring data as described in Sections 75.4.3 and 75.4.4. The monitoring data used in the analyses were selected on the basis that they represented appropriate distributions both in terms of using vulnerable drinking water sources and in providing profiles of residue levels with time. The estimated risk from a 24-h acute drinking water exposure analysis (99.9th percentile) is 27% of the aRfD for infants (1 year), the most highly exposed subpopulation (Table 75.4).

The daily aggregate assessment (food plus water) shows acceptable risk at the 99.9th percentile for the most highly exposed populations of children aged 1–2 and infants (Table 75.5). The results illustrate the impact of a probabilistic analysis by comparing the difference in the aggregate risk to the summation of the food-only and water-only population exposures at the 99.9th percentile. Aggregate models such as CARES operate by constructing consistent spatial, temporal, and demographic profiles of individuals within the population across all pathways (food, water, and

Chapter | 75  Aldicarb: Toxicity, Exposure and Risks to Humans

1629

Table 75.4  Daily Dietary Drinking Water Risk Assessment Population

Percentiles (weighted per capita) 95th

99th

99.9th

Exposure (mg/kg bw)

Percent aRfDa

Exposure (mg/kg bw)

Percent aRfDa

Exposure (mg/kg bw)

Percent aRfDa

Infants (1 year)

1.2E-05

1

9.2E-05

7

3.6E-04

27

Children (1–2 years)

5.9E-06

1

4.6E-05

4

1.7E-04

13

Adults (20–49 years)

4.4E-06

1

3.0E-05

2

9.2E-05

7

a

Percent aRfD calculated based on RfD  0.0013 mg/kg (BMDL10  0.013 mg/kg; UF  10).

residential). Thus the low probability that the individual at the high end of the exposure profile for food consumption is the same individual at the high end of the exposure profile for water consumption is appropriately assessed.

Table 75.5  Daily Aggregate Risk Assessment Population

Food

Drinking water

Aggregate

Infants (1 year)

18

27

33

Children (1–2 years)

37

13

38

Adults (20–49 yars)

13

7

14

75.5.4  Further Risk Characterization The preceding acute assessment based on total daily expos­ ures demonstrates that, using an extensive database of hazard and exposure information, aldicarb clearly meets the standard of safety typically applied to pesticides, that is, reasonable certainty of no harm. Although more information is available on aldicarb than is generally available for use in a pesticide risk assessment, further refinements to the risk assessment are still possible. Conservative assumptions have also been made in the selection and application of both the hazard and exposure data, which are protective of human health. A key refinement that could be made to the assessment is moving from the assumption that an individual consumes a complete daily dose of aldicarb from all possible sources at one time to a within-day assessment using the known times of consumption (Table 75.2). Given the rapid reversibility of aldicarb ChEI, the total daily consumption data clearly result in an overestimate of risk. Consumption data and modeling tools are now available to the risk assessor to examine the potential magnitude of this affect. Preliminary indications are that, for infants and children that ingest water at multiple times during the day, the actual risk from this source of exposure is significantly less than is determined using a composite daily exposure. Although monitoring data were used to characterize exposure, the distributions used in the risk assessment for both food and water were not the complete distributions but were biased toward high-end exposure detects. In the Special Study on Aldicarb in Potatoes performed by PDP, the majority (85%) of the composite samples were taken from states in which the product is registered. The single servings that made up those composites and were used in

Percent aRfD at the 99.9th percentile (weighted per capita)

the risk assessment for single-serve consumption items were further analyzed only when there were detectable residues in the composite (16 out of 342 composites), thus not reflecting a national distribution that includes all untreated potatoes or potatoes that were treated that had no detectable residues. The surface water monitoring study targeted community water systems with the highest aldicarb use intensity. It was an extensive study, designed to sample sources in the highest use areas with frequent sampling around the time of application (weekly) to account for levels close to peak values. Groundwater monitoring targeted the most vulnerable subset of rural wells in high aldicarb use areas except in Florida, where the state monitoring data was sufficiently extensive to be used directly. Further, the aldicarb equivalents approach to total residues in both food and drinking water typically overestimates toxicity, as residues of aldicarb sulfone, which is a significantly less potent ChE inhibitor (25), have been included, assuming tox­ icity equivalent to aldicarb. Future refinements should also incorporate updated dietary consumption data. More current nationwide dietary intake data (USDA and DHSS, 2006) and water consumption data from the National Health and Nutrition Examination

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1630

Survey (NHANES) have recently been released but have not yet been incorporated into the dietary models.

Conclusion Aldicarb has been registered for use in the United States as a systemic insecticide and nematicide on agricultural crops since 1970. The hazard and exposure profiles of aldicarb are well characterized and result in a high level of confidence in the risk assessments. Across the extensive toxicology database, ChEI is identified as the most sensitive endpoint and is utilized as the POD. The aRfD is derived from the BMDL10 of 0.013 mg/kg for RBC ChEI in a human study with an intraspecies uncertainty factor of 10. The interspecies uncertainty factor is 1, as the point of departure is based on a human study. The results of the human study provide dose–response data near the benchmark response (BMR) of 10% up to a level causing clear inhibition of RBC ChE. Utilizing a BMD approach to establish the POD avoids difficulties inherent in the NOAEL approach. The use of the lower limit of the 95% confidence interval (i.e., BMDL10) provides a more conservative POD than does the central BMD10 value and assures with high confidence that the BMR will not be exceeded. As the BMD10 and BMDL10 derived for the human study are above the lowest dose tested, the level of confidence in the values is increased further. The resultant aRfD of 0.0013 mg/kg is supported by the toxicology database for aldicarb. The pharmacokinetics of ChEI by aldicarb has been studied in both the rat and human and subjected to a rigorous statistical assessment. Peak inhibition occurs within 1 h of exposure and the rate of recovery is rapid, is doseindependent and can be adequately described by a firstorder rate constant. The rate of ChE recovery reflects the detoxification and removal of active carbamate from the body and thus can be used as a measure of residual exposure (biologically effective dose) at any time following peak inhibition. The residual expsoure with time can be calculated from the first-order rate constant and can be used to refine the risk assessments. The potential dietary risk posed by aldicarb has been assessed using extensive databases for aldicarb residues detected in food and water sources along with total daily food and water consumption data compared to the aRfD of 0.0013 mg/kg. For all scenarios examined and assuming exposure at the 99.9th percentile, food, drinking water, and aggregate analyses are acceptable and confirm a reasonable certainty of no harm. For food uses alone the most highly exposed subpopulation of children aged 1–2 utilized 37% of the aRfD. The worst-case groundwater scenario results in 27% of the aRfD utilized for infants, the subpopulation with the highest exposure from water. Surface water exposure resulted in lower risk estimates. Daily aggregate analysis, combining food and water exposures, results in

an acceptable risk of 38% of the aRfD for children aged 1–2, overall the most highly exposed population. A full characterization of the risk confirms that these values are very conservative. For aldicarb, the rapid reversibility of RBC ChEI indicates that a within-day approach to risk assessment is appropriate and would demonstrate that the percentage of aRfD consumed is lower than these estimates. In conclusion, aldicarb has a very extensive database of hazard and exposure information that, when integrated into a conservative risk assessment, clearly demonstrates that there is reasonable certainty of no harm from current aldicarb uses.

References Barraj, L., Scrafford, C., Lantz, J., Daniels, C., and Mihlan, G. (2008). Within-day drinking water consumption patterns: Results from a drinking water consumption survey. J. Expo. Sci. Env. Epid. advance online publication, 14 May 2008; doi:10.1038/jes.2008.28 pp. 1–14. Brooks, W., and Broxup, B. (1994). An Acute Study of the Time Course of Cholinesterase Inhibition by Aldicarb Technical, Aldicarb Sulfoxide, and Aldicarb Sulfone in the Rat. Unpublished Report No. 97352. Bio-Research Laboratories, Ltd. Brooks, W., and Broxup, B. (1995a). An Acute Study of the Potential Effects of Orally Administered Aldicarb Sulfoxide on Behavior and Neuromorphology in Rats. Unpublished Report No. 97350. BioResearch Laboratories, Ltd. Brooks, W., and Broxup, B. (1995b). An Acute Study of the Potential Effects of Orally Administered Aldicarb Sulfone on Behavior, Neurochemistry and Neuromorphology in Rats. Unpublished Report No. 97351. Bio-Research Laboratories, Ltd. Cimino, M. C., Galloway, S. M., and Ivett, J. L. (1984). Mutagenicity Evaluation of Aldicarb (93.47%) in the Mouse Bone Marrow Cytogenetic Assay. Unpublished Report No. 7198-8857. Litton Bionetics. Clayton, C. A., Starr, T. B., Sielken, R. L., Williams, R. L., Pontal, P. -G., and Tobia, A. J. (2003). Using a non-linear mixed effects model to characterize cholinesterase activity in rats exposed to aldicarb. J. Agr. Biol. Envir. St. 8, 420–437. Crump, K. S. (1984). A new method for determining allowable daily intakes. Fund. Appl. Toxicol. 4, 854–871. Exponent, Inc. Dietary Exposure Evaluation Model (DEEM™) http:// www.exponent.com/deem_software/. Godek, E. G., Dolak, M. C., Naismith, R. W., and Matthews, R. J. (1980a). Ames Salmonella Microsome Plate Test. TEMIK aldicarb pesticide. Unpublished Report No. PH 301-UC-004-80. Pharmakon Laboratories. Godek, E. G., Dolak, M. D., Naismith, R. W., and Matthews, R. J. (1980b). Ames Salmonella Microsome Plate Test. Aldicarb sulfone. Unpublished Report No. PH 301-UC-003-80. Pharmakon Laboratories. Godek, E. G., Dolak, M. C., Naismith, R. W., and Matthews, R. J. (1980c). Ames Salmonella Microsome Plate Test. Aldicarb sulfoxide. Unpublished Report No. PH 301-UC-002-80. Pharmakon Laboratories. Godek, E. G., Naismith, R. W., and Matthews, R. J. (1984a). Rat Hepatocyte Primary Culture/DNA Repair Test. Aldicarb technical. Unpublished Report No. PH 311-UC-005-83. Pharmakon Research International Inc. Godek, E. G., Naismith, R. W., and Matthews, R. J. (1984b). Rat Hepatocyte Primary Culture/DNA Repair Test. Aldoxycarb technical.

Chapter | 75  Aldicarb: Toxicity, Exposure and Risks to Humans

Unpublished Report No. PH 311-UC-006-83. Pharmakon Research International Inc. Hamada, N. N. (1988). One-Year Chronic Oral Toxicity Study in Beagle Dogs with Aldicarb Technical. Unpublished Report No. HLA 400706 Hazleton Laboratories America Inc. International Life Sciences Institute (ILSI). Cumulative and Aggregate Risk Evaluation System Version 3.0 (CARES®), http://cares.ilsi.org/. Ivett, J. L. (1990). Mutagenicity Test on Aldicarb Technical in the Mouse Bone Marrow Cytogenetic Assay. Unpublished Report No. HLA 12010-0-451. Hazleton Laboratories America, Inc. Jones, R. L. (2005). Comparison of regulatory estimates of drinking water concentrations with monitoring data. J. Agr. Food Chem. 53, 8835–8839. Jones, R. L., and Estes, T. L. (1995). Summary of aldicarb monitoring and research programs in the United States. J. Contam. Hydrol. 18, 107–140. Jones, R. L., and Allen, R. (2007). Summary of potable well monitoring conducted for aldicarb and its metabolites in the United States in 2005. Environ. Toxicol. Chem. 26, 1355–1360. Kelly, I., Lantz, J., Mihlan, G. J., and Young, B. M. (2006). CAREScompatible Dietary Minute Module (DMM) Version 2.2. Build 7.8.9 (3/14/2006) Technical Manual. ILSI CARES website pp. 26. Lemen, J. K. (1991). Two-Generation Reproduction Study in Rats with Aldicarb. Unpublished Report No. 656-157. Hazleton Washington, Inc. Leng, J. M., Schardein, J. L., and Blair, M. (1983). Aldicarb. Teratology study in rabbits. Unpublished Report No. 369-107. International Research and Development Corporation. Mortensen, S. R., Hooper, M. J., and Padilla, S. (1998). Rat brain acetylcholinesterase activity:developmental profile and maturational sensitivity to carbamate and organophosphorus inhibitors. Toxicology 125(1), 13–19. Moser, V. C. (1999). Comparison of aldicarb and methamidaphos neurotoxicity at different ages in the rat: behavioral and biochemical parameters. Toxicol. Appl. Pharml. 157(2), 94–106. Myers, R. C., Sleslnski, R. S., and Frank, F. R. (1985). Aldicarb in DCM without Vinyl Binder VYHD Acute Toxicity and Irritancy Study. Unpublished Report No. 48-42. Bushy Run Research Center. Nachreiner, D. J., Klonne, D. R., and Frank, F. R. (1985). Aldicarb Solution (in DCM)LC50 Aerosol Acute Inhalation Toxicity Test. Unpublished Report No. 48-136. Bushy Run Research Center. National Cancer Institute (NCI). (1979). Bioassay of Aldicarb for Possible Carcinogenicity. US DHEW Pub.No. (NIH) 79-1391. Nycum, J. S., and Carpenter, C. P. (1968). Toxicity studies on Temik and related Carbamates. Unpublished Report No. 31-48. Mellon Institute. Robinson, K., Brooks, W., and Broxup, B. (1994). An Acute Toxicity Study of the Potential Effects of Orally Administered Aldicarb Technical on Behavior and Neuromorphology in Rats. Unpublished Report No. 97235. Bio-Research Laboratories, Ltd. Robinson, K., Brooks, W., and Broxup, B. (1995). A 13-Week Study of the Potential Effects of Orally Administered Aldicarb Technical on Behavior. Unpublished Report No. 97234. Bio-Research Laboratories, Ltd. SanSebastian, J. R., Naismith, R. W., and Matthews, R. J. (1984). In Vitro Chromosome Aberration Analysis in Chinese Hamster Ovary Cells (CHO). Aldoxycarb technical. Unpublished Report No. PH 320-UC005-83. Pharmakon Research International Inc. Stankowski, L. F., Naismith, R. W., and Matthews, R. J. (1985a). CHO/ HGPRT Mammalian Cell Forward Gene Mutation Assay. Aldicarb. Unpublished Report No. PH 314-OC-003-84. Pharmakon Research International, Inc.

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Stankowski, L. F., Naismith, R. W., and Matthews, R. J. (1985b). CHO/ HGPRT Mammalian Cell Forward Gene Mutation Assay. Aldoxycarb. Unpublished Report No. PH 314-UC-002-84. Pharmakon Research International, Inc. Striegel, J. A., and Carpenter, C. P. (1962). Range Finding Test on Compound 21149. Unpublished Report No. 25-53. Mellon Institute. The Lifeline Group, Inc. LifeLine™ Version 5.0 http://www.thelifelinegroup.org/lifeline/index.htm Trutter, J. A. (1993). Combined Chronic Toxicity and Oncogenicity Study in Rats with Aldicarb Technical. Unpublished Report No. HWA 656151. Hazleton Washington, Inc. Tyl, R. W., and Neeper-Bradley, T. (1988). Developmental Toxicity Evaluation of Aldicarb Administered by Gavage to CD® (Sprague-Dawley) Rats. Unpublished Report No. 51-551. Bushy Run Research Center. U.S. Department of Agriculture (USDA). (2000). Agricultural Research Service Continuing Survey of Food Intakes by Individuals 1994-96, 1998. CD-ROM. National Technical Information Service Accession No. PB2000-500027. http://www.ars.usda.gov/Services/docs. htm?docid14531. U.S. Department of Agriculture (USDA), Agricultural Research Service, Beltsville Human Nutrition Research Center, Food Surveys Research Group and U.S. Department of Health and Human Services (DHSS), Centers for Disease Control and Prevention, National Center for Health Statistics. (2006). What We Eat in America, NHANES 20052006 Data. Available from: http://www.cdc.gov/nchs/nhanes.htm. U.S. Environmental Protection Agency (U.S. EPA), Office of Pesticide Programs and U.S. Department of Agriculture (USDA), Agricultural Research Service. (2000). Food Commodity Intake Database (FCID) Version 2.1. CD-ROM. National Technical Information Service, Accession No. PB2000-500101. http://www.ars.usda.gov/Services/ docs.htm?docid14514. U.S. Environmental Protection Agency (U.S. EPA) Office of Pesticide Programs (2000). “The Use of Data on Cholinesterase Inhibition for Risk Assessments of Organophosphorous and Carbamate Pesticides,”. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA). (2004a). 2004 Edition of the Drinking Water Standards and Health Advisories. EPA 822-R04-005, Office of Water, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA). (2004b). FIFRA Scientific Advisory Panel Meeting: A Model Comparison: Dietary and Aggregate Exposure in Calendex, CARES, and Lifeline. April 29-30, 2004. U.S. Environmental Protection Agency (U.S. EPA). (2006). EPA Human Studies Review Board Meeting Report (EPA-HSRB-06-01). U.S. Environmental Protection Agency (U.S. EPA). (2007). Human Health Risk Assessment Aldicarb. DP Barcode No. D336910. Weil, C. S., and Carpenter, C. P. (1972). Aldicarb (A), Aldicarb sulfoxide (ASO), Aldicarb sulfone (AS02) and a 1:1 mixture of ASO:AS02. Two-year feeding in the diet of rats. Unpublished Report No. 35-82. Mellon Institute. Weil, C. S., and Carpenter, C. P. (1974). Aldicarb, 18-month feeding in the diet of mice, study II. Unpublished Report No. 37-98. Mellon Institute. Weiler, M. S. (1995). Developmental Neurotoxicity Study with Aldicarb in Rats. Unpublished Report No. HWI 6224-213. Hazleton Wisconsin, Inc. West, J. S., and Carpenter, C. P. (1966). Miscellaneous Acute Toxicity Data. Unpublished Report No. 28-140. Mellon Institute. Williams, R. L., Mihlan, G. J., and Tobia, A. J. (2008). Modeling cholinesterase activity for human dietary risk assessment of carbamates insecticides. Risk Anal. 28(4), 1069–1079.

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Williams, S. D., and Harris, R. M. (1996). Nutrient, Sediment, and Pesticide Data Collected at Four Small Agricultural Basins in the Beaver Creek Watershed, West Tennessee, 1990–1995. U.S. Geological Survey Open File Report 96–366. Memphis, Tennessee, 115 pages. World Health Organization (WHO). (1999). FAO/WHO Joint Meeting on Pesticide Residues. Report of the 1998 FAO/WHO Joint Meeting

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on Pesticide Residues. Food and Agricultural Organization-United Nations. Rome, Italy. Wyld, P. J., Watson, C. E., Nimmo, W. S., and Watson, N. (1992). A Safety and Tolerability Study of Aldicarb at Various Dose Levels in Healthy Male and Female Volunteers. Unpublished Report No. 7786. Inveresk Clinical Research.

Section XI

Pyrethrins and Pyrethroids

(c) 2011 Elsevier Inc. All Rights Reserved.

Chapter 76

Pyrethroid Chemistry and Metabolism Hideo Kaneko Sumitomo Chemical Company, Ltd., Osaka, Japan

76.1  Introduction Many synthetic pyrethroid insecticides, which comprise one of the major insecticide groups, are used worldwide for controlling indoor pests and agricultural pests. Their origin or mother compounds are natural pyrethrins, the insecticidal ingredient occurring in the flowers of Tanacetum cinerariaefolium (also known as Chrysanthemum cinerariaefolium or Pyrethrum cinerariaefolium), which have been used widely for human and animal health protection by controlling indoor pest insects such as cockroaches, houseflies, and mosquitoes since ancient times. Natural pyrethrins consist of six compounds (pyrethrin I and II, jasmolin I and II, and cinerin I and II). The investigation of the chemical structures of natural pyrethrins was started in the 1920s and their absolute stereochemistry was completed and elucidated in the early 1970s (Chamberlain et al., 1998). Along with the investigations, extensive efforts on modification of chemical structures have been made in many laboratories to improve chemical properties in terms of stability in the environment (air, light, and heat) as well as better biological performance (higher selective toxicity). Synthetic pyrethroids can be classified into the so-called first- and second-generation pyrethroids. The characteristic feature of the first-generation pyrethroids, which are esters of chrysanthemic acid derivatives and alcohols having furan ring and terminal side chain moieties, is to be highly sensitive to light, air, and temperature. Therefore, these pyrethroids have been used mainly for control of indoor pests. A new chrysanthemic acid derivative has been developed: norchrysanthemic acid (a chrysanthemic acid derivative with deletion of a methyl group) (Ujihara et al., 2004). This modification leads to higher insecticidal activity and volatility compared with those of the original pyrethroid with chrysanthemic acid. On the other hand, the second-generation pyrethroids, which commonly have 3-phenoxybenzyl alcohol derivatives in the alcohol moiety, have excellent insecticidal activity as well as sufficient stability in outdoor conditions by replacement of photolabile moieties with dichlorovinyl, dibromovinyl substituent, and Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

aromatic rings. Thus, the second-generation pyrethroids have been used worldwide for agricultural pests. The previous edition of this chapter was prepared on the basis of metabolism data in the public domain approximately 10 years ago. During the past decade, only a few new pyrethroid insecticides have been introduced into the market. On the other hand, during the same time, metabolic reactions of pyrethroid insecticides have been elucidated at enzyme levels and great progress has been made regarding molecular biology and analytical instruments. For example, tissue microsomes (mainly liver) of several laboratory animals including humans, and human and laboratory animal CYP (P-450) isoforms and carboxylesterases genetically expressed in yeast, insect cells, or mammalian cell lines are now commercially available. Therefore, species differences in metabolism between humans and laboratory animals can be quantitatively and qualitatively investigated in detail and more precise risk assessment of pyrethroids can be made. Many in vivo and in vitro metabolism studies of synthetic pyrethroid insecticides including their chiral and geometrical isomers have been carried out in mammals for safety assessment. However, detailed metabolism data have not been published. In these cases, the reports of the joint Food and Agricultural Organization/World Health Organization (FAO/WHO) expert meetings on pesticide residues and of the International Programme on Chemical Safety (IPCS), Environmental Health Criteria (WHO), were referred to. This chapter discusses the metabolism of more than 20 pyrethroid insecticides in laboratory animals and humans in alphabetical order, focusing mainly on in vivo metabolism. Data on nomenclature and physical chemistry are mainly cited from A World Compendium, The Pesticide Manual, 14th edition (Tomlin, 2006), and Metabolic Pathways of Agrochemicals: Part 2. Insecticides and Fungicides (Roberts and Hutson, 1999). Following the discussion of the chemistry and metabolism of each pyrethroid, species differences, phase I and phase II enzymes responsible for the major metabolic reactions, and biomonitoring of pyrethroid metabolites are summarized. 1635

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76.2  Allethrin (bioallethrin,   d-allethrin, s-bioallethrin) Chemical name (RS)-3-Allyl-2-methyl-4-oxocyclopent-2enyl (1RS)-cis-trans-2,2-dimethyl-3-(2-methylprop-1-enyl) cyclopropanecarboxylate. Synonyms Allethrin (BSI, E-ISO, JMAF, ESA) is the common name in use. The trade name is Pynamin. d-Allethrin (trade name Pynamin Forte) is an ester of (1R)-cis-­transchrysanthemic acid and (RS)-allethrolone. Bioallethrin is an ester of (1R)-trans-chrysanthemic acid and (RS)-allethrolone. S-Bioallethrin is an ester of (1R)-trans-chrysanthemic acid and (S)-allethrolone. The CAS registry numbers are 584-79-2 (allethrin, bioallethrin) and 28434-00-6 (S-bioallethrin). Physical and chemical properties (d-allethrin) The empirical formula is C19H26O3; molecular weight is 302.4. Its form is a yellow to amber viscous liquid; its specific gravity is 1.01 at 20°C; log Kow  4.96. It is practically insoluble in water, but is soluble in most organic solvents. It is decomposed by UV light and is hydrolyzed in alkaline media. Metabolism When allethrin labeled with 14C in the acid moiety or with 3H or 14C in the alcohol moiety was orally administered to rats at 1–5 mg/kg, the 14C and 3H derived from the acid and alcohol moieties were excreted into the urine (47–51%) and feces (27–29%) within 48 h after administration. Most of the metabolites excreted into the urine are ester linkage-cleaved products [chrysanthemic dicarboxylic acid (CDCA) and allethrolone] and ester linkage-retaining

products. However, the fecal metabolites are not adequately characterized (Elliott et al., 1972; IPCS, 1989). The major metabolic reactions (Figure 76.1) of allethrin are as follows: (1) hydrolysis of the ester linkage, (2) formation of the 2,3-diol from the allyl moiety, (3) hydroxylation at the methylene position of the allyl moiety, (4) ­hydroxylation at one of the gem-dimethyl groups, and (5) oxidation at the trans-methyl group of the isobutenyl moiety. In addition, epoxidation of the double bond of the acid moiety takes place in vitro in mouse liver microsomes (Class et al., 1990). There are some species differences in in vitro microsome oxidation sites of allethrin between rats and mice: rat microsomes appear to preferentially oxidize the transmethyl group of the isobutenyl moiety. On the other hand, the major oxidation sites by mouse microsomes are the trans-methyl group of the isobutenyl group, the methylene position of the allyl group, and the 7,8 double bond of the acid moiety (Class et al., 1990). S-Bioallethrin was a good substrate in vitro for rat CYP2C11, 3A1, and 3A2 and for human CYP2C8 and 2C19 (Scollon et al., 2009).

76.3  Bifenthrin Chemical name 2-Methylbiphenyl-3-ylmethyl (Z)-(1RS)cis-3-(2-chloro-3,3,3-trifluoroprop-1-enyl)-2,2-dimethylcyclopropanecarboxylate.

HOH2C

OHC COO

COO

O

A alc

O

Allethrin

COO

HOOC

HOOC

HOOC

OH

COO

COOH CDCA

COO O

A acid

CH2OH

HOOC

O Allethrolone

COO AMC

OH

AMA

HOOC COO

COO O

OH

Figure 76.1  Metabolic pathways of allethrin in animals.

O

AMB

HO

HOOC

O

A ald

O O

O

Chapter | 76  Pyrethroid Chemistry and Metabolism

1637

Synonyms  Bifenthrin (BSI, ANSI, E-ISO) is the common name in use. The trade name is Talster. Code designations include FMC 54800. The CAS registry number is 82657-04-3. Physical and chemical properties  The empirical formula is C23H22C1F3O2; molecular weight is 422.9. Its form is a viscous liquid or a crystalline, or waxy solid; its specific gravity is 1.21 at 25°C; log Kow  6. It is less soluble (0.0001 mg/l) in water and is soluble in most organic solvents. It is rather stable in natural daylight and water (pH 5–9). Metabolism  Male and female rats were treated with 14 C-bifenthrin labeled in the acid or alcohol moiety at single oral doses of 4 and 35 mg/kg. The 14C was rapidly excreted into feces and urine, and the excretion rates of the 14 C to feces and urine were 66–83% and 13–25%, respectively. Highest residues were found in the fat, with values of slightly more than 1 ppm after low-dose administration and 8 and 16 ppm in males and females, respectively, after application of the high dose. Residue levels in other organs were in most cases less than 0.2 ppm after low-dose administration and less than 1 ppm after high-dose administration (FAO/WHO, 1992). The major fecal metabolites possessed intact ester linkage hydroxylated in the acid or alcohol moiety such as hydroxymethyl-bifenthrin, 4-OH-bifenthrin, and 3- or 4-OH-hydroxymethyl bifenthrin. Ester-cleaved products

derived from mono- and dihydroxylated parent compounds were also detected. On the other hand, the majority of urinary metabolites were ester-cleaved products such as 4-OHBP acid (4-hydroxy-2-methyl-3-phenylbenzoic acid), BP acid (2-methyl-3-phenyl-benzoic acid), 4-OH-BP ­ alcohol (4-hydroxy-2-methyl-3-phenylbenzyl alcohol), dimethoxyBP acid, 4-methoxy BP acid, dimethoxy BP alcohol, BP alcohol, TFP acid [3-(2-chloro-3,3,3-trifluoro-1-propenyl)2,2-dimethyl-cyclopropanecarboxylic acid], and cis- and trans-hydroxymethyl TFP acid. The major metabolic pathways (Figure 76.2) are considered to be hydrolysis of ester linkage, oxidation at the methyl group of the acid moiety and at the 3 and 4 positions of the phenyl group, and Omethylation. The conjugation reactions are expected to take place; however, detailed information is not available (FAO/ WHO, 1992). The tissue residues were examined after oral administration of 14C-bifenthrin at 0.5 mg/kg/day for 70 days. The peak 14C concentrations on average were 9.6 ppm in fat, 1.7 ppm in skin, 0.4 ppm in liver, 0.3 ppm in kidney, 1.7 ppm in ovaries, 3.2 ppm in sciatic nerve, 0.06 ppm in whole blood, and 0.06 ppm in plasma. Analyses were extended for an additional 85 days following cessation of dosing (depuration phase). Half-lives of 51 days (fat), 50 days (skin), 19 days (liver), 28 days (kidney), and 40 days (ovaries and sciatic nerve) were estimated from 14C-depuration. Analysis

HOH2C COOCH2

COOH CF3

CI

COOCH2

CH3

CI

CF3

CF3

Bifenthrin

TFP acid

CH3

CI

Hydroxymethyl-bifenthrin

HOH2C

HOH2C

HOOC

CH3

CH3 BP acid

BP alcohol

HOH2C CH3

COOCH2

COOCH2

(OCH3)2

Dimethoxy BP alcohol

CF3

CH3

CI

OH

CF3

4sy′-OH -bifenthrin

3′- or 4′- OH -hydroxymethyl -bifenthrin

CH2OH

HOH2C CH3

OH

4′-OH -BP alcohol

HOH2C

COOH CF3

CI

CH3

(OCH3)2

Dimethoxy -BP acid Figure 76.2  Metabolic pathways of bifenthrin in animals.

COOH CI

CF3

cis - hydroxymethyl TFP acid

HOOC

OH

CH3

CI

trans-hydroxymethyl TFP acid

HOOC

HOOC CH3 4′-OH -BP acid

OH

CH3

OCH3

4′-methoxy BP acid

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of the fat revealed that the parent chemical accounted for a majority (65–85%) of the 14C residues in fat. Bifenthrin was a good substrate in vitro for rat CYP1A1, 2C6, 2C11, and 3A1 and for human CYP2C9 and 2C19 (Scollon et al., 2009).

76.4  Cycloprothrin Chemical name (RS)--Cyano-3-phenoxybenzyl (RS)-2,2dichloro-1-(4-ethoxyphenyl) cyclopropanecarboxylate. Synonyms Cycloprothrin (BSI, ISO) is the common name in use. The trade name is Cyclosaal. Code designations include GH-414 and NK-8116. The CAS registry number is 63935-38-6. Physical and chemical properties The empirical formula is C26H21Cl2NO4; molecular weight is 482.4. Its form is a yellow to brown viscous liquid; its specific gravity is 1.256 at 25°C; log Kow  4.19. It is less soluble (0.091 mg/l) in water at 25°C, but it is soluble in most organic solvents. Metabolism On single or consecutive (once a day for 7 days) oral administration of cycloprothrin labeled with 14 C in the acid moiety to male rats at 50 mg/kg/day, the 14C was rapidly and almost completely eliminated into urine (36%) and feces (63%) within 7 days after administration. 14 C tissue residue levels reached maximum 3 h after single oral administration and thereafter decreased with time. 14C tissue residue levels after repeated administration were approximately 3.6 times higher compared with those of a

CI CI HOC2H4O

COOCH CN

76.5  Cyfluthrin (-Cyfluthrin) Chemical name (RS)--Cyano-4-fluoro-3-phenoxybenzyl (1RS)-cis-trans-3-(2,2-dichlorovinyl)-2,2-dimethylcyclopropanecarboxylate. Synonyms Cyfluthrin (BSI, E-ISO, BAN) is the common name in use. Trade names are Baythroid, Baygon aerosol, and Bayofly. Code designations include Bay FCR 1272. The CAS registry number is 68359-37-5. -Cyfluthrin is a mixture of the isomers at some rates. Physical and chemical properties The empirical formula is C22H18Cl12FNO3; molecular weight is 434.3. Its form is a colorless crystal; its specific gravity is 1.28 at 20°C; log Kow  6. It is less soluble (0.002–0.003 mg/l) in

CI CI

O

HOC2H4O- cycloprothrin

single oral dose. 14C residue levels were relatively high in the fat and skin, and 14C depletion from these tissues was slower than that from other tissues (Seguchi et al., 1991). HO-acid was a predominant metabolite in urine and feces, accounting for 39% of the dose. In addition, HOcycloprothrin, C2H5O-acid, and HOC2H4O-acid were also found in the feces and urine as minor metabolites. The major metabolic pathways of cycloprothrin (Figure 76.3) are cleavage of the ester linkage and oxidation at the ethoxy position of the acid moiety (Seguchi et al., 1991). Although metabolism of the alcohol moiety is not available, it can be predicted on the basis of metabolism of pyrethroids having the same alcohol moiety such as fenvalerate and fenpropathrin.

COOCH CN Cycloprothrin

C2H5O

O

CI CI

CI CI

CI CI

COOH

COOH

COOCH CN

HOC2H4O

HO

C2H5O HOC2H4O- acid

C2H5O- acid

CI CI COOH HO HO-acid Figure 76.3  Metabolic pathways of cycloprothrin in animals.

HO-cycloprothrin

O

Chapter | 76  Pyrethroid Chemistry and Metabolism

water at 20°C, but it is soluble in most organic solvents. It is rather stable in acidic water but unstable in alkaline water. Metabolism The acid moiety of cyfluthrin is the same as those of permethrin and cypermethrin; accordingly, metabolism of the acid moiety was not investigated because the acid moiety should undergo the same metabolic fate after ester hydrolysis. 14C derived from the alcohol moiety was rapidly and completely excreted into urine and feces after a single oral administration of 14C-alcohol-labeled preparation to rats at 0.5 and 10 mg/kg, with 55–70 and 25–35% of the dose being excreted into urine and feces, respectively. Excretion of 14C into bile was approximately 34%. The fat and sciatic nerve showed relatively higher 14C tissue residues (FAO/WHO, 1986). Major metabolic pathways (Figure 76.4) are ester cleavage and oxidation at the 4 position of the alcohol moiety. Major metabolites were 4-OH-FPB acid [4-fluoro3-(4-hydroxyphenoxy)benzoic acid] and its conjugates (glu­ curonide or sulfate), accounting for approximately 40–50% of recovered urinary 14C from rats given the labeled preparation at 0.5 mg/kg. Glycine conjugates of FPB acid (4-fluoro-3-phenoxybenzoic acid) and 4-OH-FPB acid were also found as minor metabolites. The hydroxylation at the 4 position of the alcohol moiety is as major in rats for cyfluthrin as with pyrethroids having the 3-­phenoxybenzyl alcohol or -cyano-3-phenoxybenzyl alcohol, although cyfluthrin has the fluoro atom in the 4 position of the benzyl ring (FAO/WHO, 1986). The metabolism of cyfluthrin was examined in humans after exposure of nine male volunteers to aerosol (un­labeled cyfluthrin). The cis- and trans-acid metabolites and FPB acid were detected, indicating that the ester hydrolysis occurs in humans (Leng et al., 1997). Cl

Cl

COOH DCVA

Cl

Cl COOCH

O

CN

F

O F FPB ald

Cyfluthrin O

CH2OH

O

F FPB alc

O

CHO

COOH

F FPB acid

O

CONHCH2COOH HO

F

COOH

F 4′ -OH -FPB acid

conjugates O HO

CONHCH2COOH

F

Figure 76.4  Metabolic pathways of cyfluthrin in animals.

1639

-Cyfluthrin was a good substrate in vitro for rat CYP1A1, 2C6, and 3A1 and for human CYP2C8, 2C9, and 2C19 (Scollon et al., 2009).

76.6  Cyhalothrin (-Cyhalothrin) Chemical name (RS)--Cyano-3-phenoxybenzyl (Z)-(1RS)cis-3-(2-chloro-3,3,3-trifluoropropenyl)-2,2-dimethylcyclopropanecarboxylate; -cyhalothrin, (RS)--cyano-3-phenoxy‑ benzyl (Z)-(1R)-cis-3-(2-chloro-3,3,3-trifluoropropenyl)-2,2dimethylcyclopropanecarboxylate. Synonyms Cyhalothrin (BSI, E-ISO, BAN) and cyhalothrin are the common names in use. Trade names are Cyhalon and Grenade for cyhalothrin and Karate, Warrior, and Icon for -cyhalothrin. Code designations include PP563 and ICI146814 for cyhalothrin and PP321 and ICIA0321 for -cyhalothrin. The CAS registry numbers are 68085-85-8 for cyhalothrin and 91465-08-6 for -cyhalothrin. Physical and chemical properties (cyhalothrin) The empirical formula is C23H19C1F3NO3; molecular weight is 449.9. Its form is a yellow to brown viscous oil; its specific gravity is 1.25 at 25°C; log Kow  6.9. It is less soluble (0.005 mg/l) in water at 20°C, but it is soluble in most organic solvents. It is stable to light and unstable in alkaline medium. Metabolism Cyhalothrin was rapidly excreted into urine and feces after oral administration of a 14C-labeled acid or alcohol preparation to rats at 1 or 25 mg/kg, and 14C was excreted into feces (40–65%) and into urine (20–40%) for 7 days. The fat showed the highest residue compared with other tissues. Major metabolic reactions (Figure 76.5) are ester hydrolysis and hydroxylation at the alcohol moiety. The metabolic fate of the alcohol moiety, -cyano-3-phenoxybenzyl alcohol, was the same as those of pyrethroid insecticides having the same alcohol moiety, such as fenvalerate, cypermethrin, and deltamethrin. The cyano group of the alcohol moiety of cyhalothrin is expected to undergo conversion to SCN ion. The major metabolites of the acid moiety are cyclopropylcarboxylic acid and its glucuronide, and those from the alcohol moiety are PB acid, 4-OH-Pb acid, and sulfate of 4-OH-PB acid (FAO/WHO, 1984; IPCS, 1990a). -Cyhalothrin is manufactured by crystallization of the more active pair of enantiomers from cyhalothrin. The comparative metabolism of -cyhalothrin with or without enantiomer pair A and cyhalothrin revealed that enantiomer pair A had little or no effect on the absorption, distribution tissue retention, or metabolic profiles, implying that enantiomers of cyhalothrin behave independently (IPCS, 1990a). -Cyhalothrin was a good substrate in vitro for rat CYP1A1, 2C11, 2D1, and 3A1 and for human CYP2C19 and 3A4 (Scollon et al., 2009).

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1640

Cl CF3

O

COOCH CN Cyhalothrin

glucuronide

Cl CF3

COOH

+

O

HOCH CN

CH2OH Cl CF3

COOH

O

HOOC

PB acid

glucuronide

glycine conjugate

OH HOOC

O 4′ -OH -PB acid

Sulfate Figure 76.5  Metabolic pathways of cyhalothrin in animals.

76.7  Cypermethrin (-, -, -, Cypermethrin) Chemical name (RS)--Cyano-3-phenoxybenzyl (1RS)cis-trans-3-(2,2-dichlorovinyl)-2,2-dimethylcyclopropane­ carboxylate. -Cypermethrin is a racemate comprising ((S)-(1R)-cis) and ((R)-(1S)-cis). -Cypermethrin is a mixture comprising two enantiomeric pairs in the ratio of approximately 2:3. -Cypermethrin is a mixture of enantiomers ((S)-(1R)-trans) and ((R)-(1S)-trans) in the ratio of 1:1. Cypermethrin is a mixture comprising ((S)-(1RS)-cis-trans). Synonyms Cypermethrin (BSI, ISO, ANSI, BAN) is the common name in use. Trade names are Basathrin, Cymbush, Cymperator, Cynoff, Cypersan, and several other names. Code designations include NRDC149, PP383, FMC30980, WL43467, and LE79-600. The CAS registry number is 52315-07-8. Physical and chemical properties The empirical formula is C22H19Cl2NO3; molecular weight is 416.3. Its form is a yellow-brown viscous semisolid; its specific gravity is 1.24 at 20°C; log Kow  6.6. It is less soluble (0.004 mg/l) in water at 20°C, but it is soluble in most organic solvents.

It is relatively stable to light in weakly acidic water, but it is unstable in alkaline medium. Metabolism On single oral administration of each of 14 C-(1RS)-trans- and (1RS)- cis-cypermethrin labeled in the benzyl ring, the cyclopropane ring, or the CN group to male and female rats at 1–5 mg/kg, 14C from the acid and alcohol moieties was rapidly and almost completely excreted into the urine and feces. The 14C from the CN group was relatively slowly excreted in the urine and feces, with the total recovery being 50–67%. The tissue residues of rats treated with the acid- or alcohol-labeled preparations were generally very low except for the fat (1 ppm). In contrast, the CN-labeled preparation showed relatively high residue levels, especially in the stomach (contents), intestines, and skin (Crawford et al., 1981a). The major metabolic reactions (Figure 76.6) of transand cis-cypermethrin were cleavage of ester linkage, oxidation at the trans- and cis-methyl cyclopropane ring and at the 4 position of the phenoxy group, and conversion of the CN group to SCN ion. The following minor species differences were observed: (1) oxidation at 5 and 6 positions of the alcohol moiety was observed in mice but not in rats; and (2) ester metabolites such as 2-OH-, 5-OH-, and trans-OH,4-OH-cypermethrin were detected in feces of mice but not of rats. The remarkable species difference in metabolites was the Pb acid–taurine conjugate, which was the predominant metabolite in mice, but it was not detected in rats. The ester linkage of cis-cypermethrin seems to be more stable than that of the corresponding trans isomer, based on the nature of urinary and fecal metabolites and excretion rate (Crawford et al., 1981b; Edwards et al., 1990; Hutson and Casida, 1978; Hutson et al., 1981). There are also species differences of conjugation reactions of the alcohol moiety in other species; Pb acid– ­glycine is predominant in sheep, cat, gerbil, and ferret; Pb acid–taurine in ferret; Pb acid–glycylvaline in mallard duck; and Pb acid–glucuronide and/or 4-OH-Pb acid– glucuronide in hamster, guinea pig, marmoset, and rabbit. The rat was unique in utilizing sulfuric acid for conjugation of 3-phenoxybenzyl moiety among animal species tested (Huckle et al., 1981). Metabolism of cypermethrin (cis:trans  1:1) in humans was investigated after oral administration to six male volunteers at 3.3 mg per person. The four metabolites from the acid and alcohol moieties were analyzed in urine. The amount of cis- and trans-Cl2CA was approximately equal to that of PB acid and 4-OH-PB acid. The ratio of trans- to cis-Cl2CA was on average 2:1, implying that ester hydrolysis is the major metabolic pathway and that the trans isomer was more rapidly hydrolyzed than the cis isomer, as is the case with rats. On the other hand, dermal application of cypermethrin (cis:trans  56:44) led to formation of a different ratio of metabolites (the ratio of trans- to cis-Cl2CA is 1:1.2) from oral administration. The estimated percutaneous absorption rate (1.2%) for 24 h is much less than that

Chapter | 76  Pyrethroid Chemistry and Metabolism

1641

Cl Cl

O

COOCH

HOCH CN

CN Cypermethrin Cl taurine conjugate glucuronide

Cl

glycine conjugate

Cl

Cl

Cl Cl

glucuronide

Cl Cl

CH2OH COOH

OHC

O

HOOC

CH2OH COOCH

O

Cl

Cl

sulfate

HO OH O

HO

6-OH -PB acid glucuronide

HOOC

CN

O c-OH -CI2CA -lactone

O

O

CN 4′-OH -cyper COOCH

taurine conjugate

OH HOOC

COOCH

glucuronide

O PB acid

Cl

Cl

Cl

O

glycine conjugate

CN t-OH, 4′-OH -cyper

t-CH2OH -CI2CA

Cl

glucuronide

PB alc

CH2OH COOCH

O

CN t-OH- cyper

COOH Cl2CA

HOH2C

O

O 4′-OH-PB acid

OH 2′-OH -cyper

sulfate OH

Cl Cl

COOCH

O

HOOC

O

glucuronide

CN

OH 5-OH -cyper

OH 5 -OH -PB acid

Figure 76.6  Metabolic pathways of cypermethrin in animals.

with rats (12% in vivo and 34% in vitro), indicating that the rat model may overestimate percutaneous absorption for humans (Capt et al., 2007; Woollen et al., 1992). Species differences for ester hydrolysis were examined with carboxylesterases from porcine, rabbit (mixture of carboxylesterases), human hCE1 (a major isoform in liver), hCE-2, and mouse (NM133960 and BAC36707) for the eight isomers of cypermethrin derivatives with fluorescence. It was found that all carboxylesterases consistently hydrolyzed the trans-isomers more rapidly than the corresponding cis-isomers (Huang et al., 2005; Nishi et al., 2006). Cypermethrin isomers were good substrates in vitro for rat CYP1A1, 2A1, 2C6, 2C11, 3A1, and 3A2 and for human CYP1A2, 2C8, 2C19, and 3A4 (Scollon et al., 2009).

76.8  Cyphenothrin Chemical name (RS)--Cyano-3-phenoxybenzyl (1R)-cistrans-2,2-dimethyl-3-(2-methylprop-1-enyl)cyclopropane­ carboxylate. Synonyms Cyphenothrin (BSI, E-ISO) is the common name in use. The trade name is Gokilaht. Code designations include S-2703 Forte. The CAS registry number is 39515-40-7.

Physical and chemical properties The empirical formula is C24H25NO3; molecular weight is 375.5. Its form is a viscous yellow liquid; its specific gravity is 1.08 at 25°C; log Kow  6.3. It is less soluble (0.009 mg/l) in water at 25°C, but it is soluble in most organic solvents. It is stable under normal storage conditions. Metabolism Single oral or subcutaneous administration of 14C-trans- or cis-cyphenothrin labeled in the acid or alcohol moiety to rats at 2–4 mg/kg resulted in almost complete elimination of the 14C from the animal body. Major excretion routes with the acid- or alcohol- (except for the CN group) labeled preparation were the urine and feces. The total recovery of the 14C within 7 days after administration of these labeled preparations was more than 93% in urine and feces. On the other hand, the 14C derived from the CN group was more slowly excreted. In addition, 4–6% of the 14C was expired as 14CO2. The total 14C recovery was 60–80% for the trans and cis isomers. The three labeled preparations of the trans and cis isomers showed more urinary excretion of the 14C with subcutaneous than with oral administration (Kaneko et al., 1984c). The 14C tissue residue levels 7 days after single oral or subcutaneous administration of each of the 14C-labeled preparations of trans- and cis-cyphenothrin were measured. With the acid- and alcohol-labeled preparations of the trans and cis isomers, the tissue residue levels were

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1642

generally very low. On the other hand, the CN-labeled preparation showed relatively higher tissue residues than other labeled preparations. Both trans and cis isomers underwent the following major metabolic reactions (Figure 76.7): (1) oxidation at the 2- and 4-phenoxy positions of the alcohol moiety, (2) oxidation at the isobutenyl and the gem-dimethyl groups of the acid moiety, (3) cleavage of ester linkage, (4) conversion of the CN ion to SCN ion and CO2, and (5) conjugation of the resulting carboxylic acids and phenols with glucuronic acid, sulfuric acid, and glycine. In vivo and in vitro comparative metabolism studies of phenothrin and cyphenothrin showed the following results: (1) the trans isomers of cyphenothrin and phenothrin were hydrolyzed more rapidly in vitro (liver homogenates) and in vivo than the corresponding cis isomers, and ciscyphenothrin was hydrolyzed to a larger extent than cisphenothrin, and (2) plasma esterases showed a different substrate specificity from the liver esterases and hydrolyzed the trans and cis isomers of cyphenothrin and phenothrin to nearly the same extent. From the results of the in vivo and in vitro studies, the CN group introduced into

COOH

Chemical name (S)--Cyano-3-phenoxybenzyl (1R)-cis-3(2,2-dibromovinyl)-2,2-dimethylcyclopropanecarboxylate. Synonyms Deltamethrin (BSI, E-ISO) is the common name in use. Trade names are Decis, Decasyn, Butox, KOthrine, Kordon, and Sadethrin. Code designations include NRDC 161, AEF 032640, and RU 22974. The CAS registry number is 52918-63-5. Physical and chemical properties The empirical formula is C22H19Br2NO3; molecular weight is 505.2. Its form is a colorless crystal; its specific gravity (bulk density) is 0.55 at 25°C; log Kow  4.6. It is less soluble (0.0002 mg/l)

HOOC

COOH wt-acid-c-CA

wt-acid-c-CA

76.9  Deltamethrin

HOOC

HOOC HOOC

the molecule did not affect the biodegradability of transcyphenothrin but, rather, made cis-cyphenothrin more biodegradable than cis-phenothrin. These in vivo metabolic profiles (ester hydrolysis rate and excretion pattern into urine and feces) may be mainly determined by activity and/or substrate specificity of the liver carboxylesterases (Kaneko et al., 1984c).

COOH

wt-acid-t-CA

wt-acid-t-CA

COOH

HOH2C HOH2C HOH2C

HOH2C

COOH

COOH

COOH

wc-alc-t-CA

wc-alc-c-CA

wt-acid-c-CA

wt-alc-t-CA

COOH

glucuronide c-CA

COOH

COOCH cis-Cyphenothrin

t-CA

O

O

COOCH

CN

trans-Cyphenothrin

glucuronide

COOH

CN

Ester metabolites

R4

HOCH

R1 R2 R1 1) COOH 2) CH3 3) CH2OH 4) CH3 5) CH3 6) CH3

HOCH

O

OH

CN

CN O

COOCH

CN–

CN R2 CH3 COOH CH3 CH2OH COOH COOH

O

R3 R3 H H H H H OH

R4 CH3 CH3 CH3 CH3 CH2OH CH3

OHC

O SCN–

OHC

CO2

O OH

PB ald HOOC

O

HOOC

O OH

glycine conjugate

Figure 76.7  Metabolic pathways of cyphenothrin in animals.

PB acid

glucuronide

sulfate

2′-or 4′-OH-PB acid

glucuronide

Chapter | 76  Pyrethroid Chemistry and Metabolism

1643

in water at 25°C, but it is soluble in most organic solvents. It is stable to air and in acidic conditions but rather unstable in alkaline medium. Metabolism On oral administration to rats at 0.60–1.64 mg/kg, the acid and alcohol moieties of deltamethrin were almost completely eliminated from the body within 2–4 days. On the other hand, the CN group was eliminated more slowly than the acid and alcohol moieties, with the total recovery during 8 days being 79% of the dosed radiocarbon (43 and 36% in the urine and feces, respectively). 14C tissue residues with the deltamethrin preparation labeled in the acid moiety or in the alcohol moiety were generally very low, whereas the fat showed somewhat higher residue levels (0.1–0.2 ppm). The 14C derived from the CN group showed relatively high residue levels, especially in skin and stomach. Essentially all the 14C in the stomach was SCN ion (Ruzo et al., 1978). The major metabolic reactions (Figure 76.8) of delta­ methrin in rats are oxidation at the trans-methyl relative

HOOC

glucuronide

O

Br

glucuronide

HOOC

O

Br

4′-OH-PB acid HOOC

OH

O 2′-OH-PB acid

CN–

SCN–

NH

HOHC

O

Br

OH

N COOH H ITCA

Deltamethrin

Br

O

Br

Br

PB acid HOH2C

c-Br2CA

OH

O

CN

COOCH

HOH2C

O

Br

CN t-CH2OH-dec

Br

O

COOCH

O

CN 4′-OH, t-OH-dec

OH

PB alc Br

O

Br

glucuronide 4′-OH-PB alc

OH

Figure 76.8  Metabolic pathways of deltamethrin in animals.

COOH

t-CH2OH-c-Br2CA

glucuronide

HOH2C

glycine conjugate

O

HOH2C

O

Br

glycine conjugate

COOH

HOH2C

OHC

HOOC

Br

CN 4′-OH-dec

COOCH

PB ald taurine conjugate glucuronide

Br

OH

COOCH

Br

CN

S

Br

O

CN 5-OH-dec

OH

glucuronide sulfate

COOCH

Br OH 5-OH-PB acid

sulfate

to the carbonyl group of the acid moiety and 2, 4, and 5 positions of the alcohol moiety; cleavage of ester linkage; conversion of the CN group to SCN ion and 2-iminothiazolidine-4-carboxylic acid (ITCA); and conjugation of these carboxylic acid and phenol derivatives with sulfuric acid, glycine, and/or glucuronic acid. The oxidation reactions were mediated mainly by rat CYP1A2, 1A1, 2C6, 2C11, and 3A2; these liver CYP isoforms metabolized delta­methrin more efficiently than liver or plasma carboxylesterases, although delta­ methrin was hydrolyzed effectively by them (Anand et al., 2006; Crow et al., 2007; Godin et al., 2007). Although the major metabolic pathways in mice are similar to those in rats, there are the following species differences in metabolism: (1) there are differences in the amino acid conjugation reactions of the alcohol moiety (rat, glycine; mouse, taurine), (2) rats produce more phenolic metabolites than mice, and (3) mice produce transhydroxymethyl cyclopropanecarboxylic acid to a larger extent than rats (Ruzo et al., 1978, 1979).

COOCH CN 2′-OH-dec

O

OH

sulfate glucuronide

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1644

A human metabolism study of deltamethrin was carried out in three volunteers after a single oral dose of 3 mg of 14C-deltamethrin per person. The 14C was more rapidly excreted into urine (51–59%) than into feces (10–26%); total excretion of 14C was 64–77% of the dose for 96 h (IPCS, 1990b). Deltamethrin was not hydrolyzed by human serum and its purified carboxylesterase. Deltamethrin was hydrolyzed by human CE-1 and rat hydrolase A more readily than by human CE-2 and rat hydrolase B, respectively (Godin et al., 2006). Deltamethrin was a good substrate in vitro for human CYP2C8, 2C19, and 3A5 but not for 2C9 (Godin et al., 2007) .

HOH2C COO trans, cis -Empenthrin

O HO 2-oxo-EMPA

glucuronide

76.10  Empenthrin Chemical name (E)-(RS)-1-Ethynyl-2-methylpent-2-enyl (1RS)-cis-trans-2,2-dimethyl-3-(2-methylprop-1-enyl) cyclopropanecarboxylate. Synonyms Empenthrin (BSI, ISO) is the common name in use. The trade name is Vaporthrin. Code designations include S-2852 Forte((EZ)-(1R)-isomers). The CAS registry number is 54406-48-3. Physical and chemical properties The empirical formula is C18H26O2; molecular weight is 274.4. Its form is a yellow liquid; its specific gravity is 0.927 at 20°C; log Kow  5 (estimate). It is less soluble (0.111 mg/1) in water at 25°C, but it is soluble in most organic solvents. It is stable under normal conditions. Metabolism When single oral administration of (1R)cis- or (1R)-trans-empenthrin labeled with 14C in the alcohol moiety was given to rats at 3–600 mg/kg (female), 97–106% of the dosed 14C was rapidly eliminated into urine and feces within 7 days after administration. Monitoring of the expired air indicated that less than 1.1% of the dose was excreted as 14CO2. Urinary, fecal, and exhaled 14C excretion accounted for 22–41, 60–74, and 1–2% of the dose, respectively (Isobe et al., 1992). 14 C tissue levels reached maxima at 1–8 h after administration of the cis or trans isomer at 3 mg/kg and decreased thereafter. Liver and kidney tissues showed higher 14C concentrations than other tissues. No notable sex-related difference was observed in distribution or excretion of the radioactivity. 14C tissue residues were lower in rats receiving the trans isomer than in those receiving the cis isomer. The parent compound accounted for 7–13 and 17–26% of the dose in the feces of rats receiving the cis and trans isomers, respectively. The major metabolites were 1-ethynyl-2-methylpent-2-enol (EMPA), 6-OH-EMPA, and 2-oxoEMPA and their glucuronides. An ester-retaining metabolite by E-methyl hydroxylation at the isobutenyl group in the acid moiety was also found. Major metabolic reactions (Figure 76.9) in rats were cleavage of the ester linkage and glucuronide formation of the resulting alcohol derivatives; as a minor pathway, oxidation

COO

HO

E-CH2OH

HO

EMPA

OH 6-OH-EMPA

glucuronide

glucuronide

Figure 76.9  Metabolic pathways of empenthrin in animals.

of the methylene group in the alcohol moiety and hydration of the triple bond in the alcohol moiety were found. In addition, oxidation at the methyl group of the isobutenyl group occurred (Isobe et al., 1992).

76.11  Etofenprox Chemical name 2-(4-Ethoxyphenyl)-2-methylpropyl 3phenoxybenzyl ether. Synonyms Etofenprox (ISO, BSI, INN) is the common name in use. Code designations include MTI-500. The trade name is Trebon, Vectron, and Boxer. The CAS registry number is 80844-07-01. Physical and chemical properties The empirical formula is C25H28O3; molecular weight is 376.5. Its form is a white crystal; its specific gravity is 1.17 at 20°C; log Kow  6.9. It is less soluble (0.022 mg/l) in water at 25°C, but it is soluble in most organic solvents. It is stable to light and in acidic and alkaline medium. Metabolism The metabolism of etofenprox has been studied in rats and dogs. On single oral administration of 14C-etofenprox (a 1:1 mixture of [1-14C-propyl]-etofenprox and [-14C-benzyl]etofenprox to both sexes of rats at 30 or 180 mg/kg, 14C excretion rates in feces and urine were 87–90 and 7–9%, respectively, of the dosed 14C 5 days after administration. No 14 C was found in expired air. Plasma 14C reached peak levels after 3–5 h. 14C excreted into bile was found to account for 10–30% and unchanged etofenprox was not found in the bile. 14C tissue concentration was the highest in fat, as unchanged parent compound. When 14C-etofenprox was administered to pregnant rats (gestation day 10 to day 16) at 30 mg/kg/day, the 14C was transferred to the fetus through the placenta; however, their levels were low compared to other tissues of mother animals. Etofenprox was secreted in milk as the unchanged compound (FAO/WHO, 1993). The major fecal metabolites were deethylated ­etofenprox (DE) and 4-OH-etofenprox (4OH). The major

Chapter | 76  Pyrethroid Chemistry and Metabolism

C2H5O

1645

C2H5O

O

CH2OCH2

CH2OCH2

O OH

Etofenprox

4′ OH

C2H5O

O

CH2OCO

C2H5O

CH2OH

α - CO

HO

CH2OCH2

C2H5O

O

CH2OCO

O OH

DE 4′ OH

DE

glucuronide or sulfate HO

HOOC

O

CH2OH

HOOC

O OH glucuronide or sulfate

Figure 76.10  Metabolic pathways of etofenprox in animals.

biotransformation routes (Figure 76.10) involved O-deethyl­ ation of the ethoxyphenyl moiety and hydroxylation of the phenoxybenzyl moiety followed by conjugation with ­glucuronide or sulfate. Oxidation of the -CH2 group followed by hydrolysis represents an additional route. On single oral administration of 14C-etofenprox (a 1:1 mixture of [1-14C-propyl]-etofenprox and [-14C-benzyl]etofenprox to beagle dogs of each sex at 30 mg/kg, total excretion in feces was 90 and 6% of the dosed 14C in urine over 5 days after administration. Based on the results of metabolites in excreta, the total estimated gastrointestinal absorption was 14–51%. Tissue concentrations were highest in liver. The results indicate a lower gastrointestinal absorption rate in dogs than in rats. The major biotransformation routes were the same as in rats (FAO/WHO, 1993).

76.12  Fenpropathrin Chemical name (RS)--Cyano-3-phenoxybenzyl 2,2,3,3tetramethylcyclopropanecarboxylate. Synonyms Fenpropathrin (BSI, E-ISO, ANSI) is the common name in use. Trade names are Rody, Danitol, Meothrin, Herald, and Fenprodate. Code designations include S-3206. The CAS registry number is 64257-84-7.

Physical and chemical properties The empirical formula is C22H23NO3; molecular weight is 349.4. Its form is a yellow-brown solid; its specific gravity is 1.15 at 25°C; log Kow  6. It is less soluble (0.014 mg/l) in water at 25°C, but it is soluble in most organic solvents. It is stable to light but is unstable in alkaline medium. Metabolism Single oral administration of 14C-acid- and 14 C-alcohol-labeled fenpropathrin preparations to rats at 2.4–26.8 mg/kg resulted in almost complete elimination of the 14C from the animal body within 7 days. Major excretion routes of 14C-acid and 14C-alcohol preparations were the urine and feces. The 14C recoveries with the acid and alcohol preparations were 96–102% (urine, 27–44%; feces, 58–70%) and 96–98% (urine, 26–43%; feces, 54–71%), respectively. The 14C excretion pattern into the urine and feces was very similar between both labeled preparations (Kaneko et al., 1987). With 14C-acid and -alcohol preparations, the tissue residue levels 7 days after single oral administration of each of the 14C-labeled fenpropathrin preparations were generally very low. However, the fat showed a slightly higher residue level for both labeled preparations compared with other tissues (Kaneko et al., 1987). Fenpropathrin was rapidly metabolized in rats via cleavage of the ester linkage, oxidation at the methyl group of

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1646

glucuronide TMPA

CN

CN Fenpropathrin

glucuronide

CN

CN

4′-OH,CH2OH-Fenp.

CH2OH-Fenp.

OH

HOOC

HOOC COOH TMPA-COOH

O

COOCH CH COOH-Fenp.

HOH2C CH2 glucuronide

O

COOCH

O

COOCH

COOH TMPA-CH2OH

OH

4′-OH-Fenp.

HOH2C

HOH2C

HOH2C

O

COOCH

O

COOCH

COOH

HOCH

O

CN O

CO TMPA-CH2OH-lactone

HOCH

O

OH

CN

OHC

O

OHC

O OH

PB ald HOOC

O

HOOC

O OH

glycine conjugate PB acid

sulfate

2′- or 4′ -OH -PB acid

Figure 76.11  Metabolic pathways of fenpropathrin in animals.

the acid moiety and at the 4 position of the alcohol moiety, and conjugation with sulfuric acid, glucuronic acid, and amino acid (Figure 76.11). The CN group is presumed to undergo the same metabolic reaction – ­ conversion of the CN group to SCN ion – as those observed in pyrethroid insecticides having -cyano-3-phenoxybenzyl alcohol, such as cyphenothrin, deltamethrin, and fenvalerate. The major urinary metabolites were sulfate of 4-OH-PB acid from the alcohol moiety and free and glucuronides of 2,2,3,3tetramethylcyclopropanecarboxylic acid (TMPA) and its hydroxymethyl derivatives from the acid moiety. The major fecal metabolites retained ester linkage with oxidation at the 4 position of the alcohol moiety and the trans-methyl group of the acid moiety (Crawford and Hutson, 1977; Kaneko et al., 1987).

76.13  Fenvalerate (Esfenvalerate) Chemical name (RS)--Cyano-3-phenoxybenzyl (RS)-2-(4chlorophenyl)-3-methylbutyrate; esfenvalerate is (S)--cyano3-phenoxybenzy (S)-2-(4-chlorophenyl)-3-methylbutyrate. Synonyms Fenvalerate (BSI, E-ISO, ESA) is the common name in use. Trade names are Arfen, Devifen,

Sumicidin, Pydrin, and several other names. Code designations include S-5602 and WL43775. Esfenvalerate (BSI, ISO) is an insecticidally active isomer of four isomers of fenvalerate and is the common name in use. Trade names are Sumi-alpha and Asana. Code designations include S5602A, DPX-YB656, and S-1844. The CAS registry numbers are 51630-58-1 for fenvalerate and 66230-04-4 for esfenvalerate. Physical and chemical properties (fenvalerate) The empirical formula is C25H22C1NO3; molecular weight is 419.9. Its form is a viscous yellow or brown liquid and sometimes partly crystalline at room temperature; its specific gravity is 1.17 at 25°C; log Kow  6.2. It is less soluble (0.01 mg/1) in water at 25°C, but it is readily soluble in most organic solvents. It is relatively stable in acidic media but unstable in alkaline medium. Metabolism Single oral administration of the 14C preparations of fenvalerate and its (2S) isomer labeled in the acid or alcohol moiety to both sexes of rats and mice at 6.7–8.4 mg/kg resulted in almost complete elimination of the 14C from the animal body. Major excretion routes for the 14C in both animal species were the urine and feces. The total recovery of the 14C 6 or 7 days after administration was 93–102% in rats and mice. In contrast, the 14C from the

Chapter | 76  Pyrethroid Chemistry and Metabolism

O CO Cl-B acid-lactone

Cl

CH2

CH2

CH2 Cl

O OH CO 2,3-OH-CPIA-lactone

Cl

CO 3-OH -CPIA-lactone

Cl

Cl

Cl

COOH Cl-B acid

COOH Cl

OH COOH 2,3-OH-CPIA

conjugate

Cl

COOCH

CO Cl-BD acid-anhydride

HOCH

CPIA

Cl

O

SCN-

O

OHC

CO2

O

COOCH CN

OH

2′- or 4′ -OH -Fenvalerate O

HOCH

CN

CN-

glucuronide

COOH

CN Fenvalerate

CO Cl

Cl

COOH 3-OH-CPIA

COOH

O

O

CH2OH

CH2OH

CH2OH

1647

OH

CN O

O

OHC

OH

PB ald HOCH2

O

PB alc

HOCH2 HOOC

O PB acid

taurine conjugote

O

HOOC

OH

O OH 4′ -OH -PB alc

2′ -or 4′ -OH -PB acid

glycine conjugate

glucuronide

sulfate

glucuronide

Figure 76.12  Metabolic pathways of fenvalerate in animals.

14

CN preparation of fenvalerate and its (2S) isomer was more slowly excreted than other 14C preparations and mainly into the urine and feces. In addition, approximately 6–14% of the 14C was expired as 14CO2 in the two species. The total recovery of the 14C was 75–81% in rats and 88–89% in mice (Kaneko et al., 1981a; Ohkawa et al., 1979). Single oral administration of 14C-acid and 14C-alcohol fenvalerate to beagle dogs resulted in rapid 14C elimination from the animal bodies. Major 14C excretion routes were the urine and feces. The 14C recovery was 87% (55.5 and 31.6% in the feces and urine) and 79% (42.3 and 36.8% in the feces and urine) 3 days after oral administration of the acid- and alcohol-labeled preparations, respectively (Kaneko et al., 1984a). The 14C tissue residue levels 6 or 7 days after administration of 14C-labeled preparations to both sexes of rats and mice were determined. With the preparations of fenvalerate and its (2S) isomer labeled in the acid and alcohol moieties except for the CN group, the residue level in the fat was relatively higher in rats and mice, whereas the residue levels in other tissues, including blood, hair, liver, kidney, and skin, were low. However, administration of the CN-labeled preparations resulted in somewhat higher tissue residues, in general, compared with other labeled preparations. Higher residues were especially found in the hair, skin, stomach, blood, and fat, and it was found that most of these residues

were due to retention of SCN ion. The 14C levels in these tissues were lower in mice than in rats. Fenvalerate underwent the following major metabolic reactions (Figure 76.12): hydroxylation at the 4-phenoxy position of the alcohol moiety and the C2 and C3 positions of the acid moiety, cleavage of the ester linkage, conversion of the CN group to SCN ion and CO2, and conjugation of the resulting carboxylic acids, phenols, and alcohols with glucuronic acid, sulfuric acid, and/or glycine. With the alcohol moiety, the following apparent species differences were observed between dogs and rodents such as rats and mice: (1) hydroxylation at both the 2 and 4 positions of the alcohol moiety occurred in rats and mice but only at the 4 position in dogs, (2) PB alc and 4-OHPB alc from the alcohol moiety were obtained from dogs to a considerable extent but were not detected in rats or mice, and (3) PB acid-glycine was found to a larger extent in dogs than in rats or mice. There were also the following remarkable species differences, particularly in the major conjugates of the alcohol moiety: Pb acid–glycine was predominant in dogs, 4-OH-Pb acid–sulfate in rats, and Pb acid–taurine in mice (Kaneko et al., 1981a; Ohkawa et al., 1979). A comparative metabolism study of the four optical isomers of fenvalerate was carried out. The 14C-labeled preparations of the four isomers labeled in the acid ­moiety

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1648

were administered to rats and mice; out of the four isomers, only the (2R, S) isomer produced cholesterol ester conjugate, which is an ester of the acid moiety (2(4-chlorophenyl)isovaleric acid (CPIA)) of the (2R, S) isomer and cholesterol. This metabolite was found in relatively larger amounts in spleen, lymph node, adrenal, and liver tissues than other tissues. This conjugate was demonstrated to be formed by transesterification reaction, not by any of three known pathways of cholesterol ester biosynthesis [acyl-CoA, cholesterol o-acyltransferase (ACAT); lecithin, cholesterol o-acyltransferase (LCAT); and cholesterol esterase], and to be a causative agent of granulomatous changes that resulted from long-term or subacute administration of fenvalerate but not esfenvalerate (Kaneko et al., 1986a, 1988; Miyamoto et al., 1986; Okuno et al., 1986). In addition, a comparative metabolism study of fenvalerate and esfenvalerate was carried out, and the results showed that there were no significant differences in metabolism between fenvalerate and esfenvalerate except for formation of a cholesterol ester conjugate from fenvalerate and that the other three isomers of fenvalerate did not seem to affect the absorption, excretion, distribution (including placental transfer), and biotransformation of esfenvalerate (Isobe et al., 1990; Shiba et al., 1990). Ester hydrolysis of the four isomers of fenvalerate derivatives with fluorescence was examined with carboxyl­ esterases of porcine, rabbit, human (hCE1), and mouse (NM133960 and BAC36707). It was found that all carboxyesterases consistently hydrolyzed (2R, S) isomers more rapidly among the four isomers and that the chiral center of the acid moiety more greatly influenced stereoselective ester hydrolysis than that at the position of the alcohol moiety (Huang et al., 2005). Esfenvalerate was hydrolyzed

HO

COOCH

O

CF2HO

COOCH

effectively by rat serum and its purified carboxylesterase, but it was not hydrolyzed by human serum and its purified carboxylesterase. Esfenvalerate was hydrolyzed by human CE-1 (a major isomer in liver) and rat hydrolase A more readily than by human CE-2 and rat hydrolase B, respectively (Godin et al., 2006, 2007). Esfenvalerate was a good substrate in vitro for rat CYP1A1, 2C6, 2C11, and 3A2 and for human CYP2C8, 2C9, 2C19, and 3A5 (Godin et al., 2007).

76.14  Flucythrinate Chemical name (RS)--Cyano-3-phenoxybenzyl (S)-2-(4difluoromethoxyphenyl)-3-methylbutyrate. Synonyms Flucythrinate (BSI, E-ISO, ANSI) is the common name in use. Trade names are Cybolt, Fluent, and Pay-Off. Code designations include AC 222705 and CL222705. The CAS registry number is 70124-77-5. Physical and chemical properties The empirical formula is C26H23F2NO4; molecular weight is 451.4. Its form is a dark amber viscous liquid; its specific gravity is 1.19 at 22°C; log Kow  4.7. It is less soluble (0.5 mg/l) in water at 21°C, but it is soluble in most organic solvents. It is unstable in alkaline conditions. Metabolism Flucythrinate is the same as fenvalerate in terms of chemical structure except for substitution of the benzene ring of the acid moiety: chlorine atom for fenvalerate and the difluoromethoxy group for flucythrinate. When 14C preparations of flucythrinate labeled in the acid or alcohol moiety were administered to rats at 19.7 mg/kg, 14 C was excreted into urine (20–30%) and feces (70–73%). The major metabolic reactions (Figure 76.13) are cleavage of ester linkage and oxidation at the gem-dimethyl

O

CF2HO

CN Flucythrinate

CN

CF2HO

CONH2

metabolites from the alcohol moiety (please refer to fenvalerate)

COOH

CH2OH HO

COOH

CF2HO

COOH

CH2 CF2HO

Figure 76.13  Metabolic pathways of flucythrinate in animals.

CO

O

COOCH

CF2HO

CONHCH2COOH

O

Chapter | 76  Pyrethroid Chemistry and Metabolism

1649

groups of the acid moiety and the 4’ position of the alcohol moiety. In addition, de-difluoromethylation takes place. The major metabolites from the alcohol moiety are the same as those from fenvalerate, and the major metabolite from the acid moiety is 2-(4-difluoromethoxyphenyl)-3methylbutyric acid (FAO/WHO, 1985).

76.15  Flumethrin Chemical name (RS)--Cyano-4-fluoro-3-phenoxybenzyl 3-(,4-dichlorostyryl)-2,2-dimethylcyclopropanecarboxylate. Synonyms Flumethrin (BAN) is the common name in use. Trade names are Bayticol and Bayvarol. Code designations include BAY VI6045 and BAY Vq1950. The CAS registry number is 69770-45-2. Physical and chemical properties The empirical formula is C28H22Cl2FNO3; molecular weight is 510.4. Its form is a yellowish and highly viscous oil; log Kow  6.2. It is less soluble (0.001 mg/l) in water at 20°C, but it is soluble in most organic solvents. Metabolism When 14C-flumethrin labeled in the acid moiety was administered to rats at 1–5 mg/kg, absorption was rapid but incomplete. The maximum 14C levels in plasma were obtained approximately 8 h after administration. The major urinary metabolites are 3-(2-chloro-2-(4chlorophenyl)ethenyl)-2,2-dimethylcyclopropanecarboxylic acid (flumethrin acid) and its glucuronide from the acid moiety, and 4-fluoro-3-phenoxybenzoic acid (FPB acid), 4-OH-FPB acid, and their glycine conjugates from the alcohol moiety. The major fecal metabolites are unchanged flumethrin and flumethrin acid.

CI

C CH

COOCH

CI

CN

The major metabolic reactions (Figure 76.14) of flu­ methrin are ester hydrolysis and oxidation of the 4’ position of the alcohol moiety (FAO/WHO, 1996).

76.16  t-Fluvalinate (Fluvalinate) Chemicalname(RS)--Cyano-3-phenoxybenzyl N-(2-chloro,,-trifluoro-p-tolyl)-d-valinate; fluvalinate is (RS)-cyano-3-phenoxybenzy N-(2-chloro-,,-trifluoro-p-tolyl)dl-valine. Synonyms t-Fluvalinate (BSI, E-ISO) is the common name in use. Trade names are Mavrik, Aquaflow, and Klartan. Code designations include SAN5271. The CAS registry numbers are 102851-06-9 for t-fluvalinate and 69409-94-5 for fluvalinate. Physical and chemical properties The empirical formula is C26H22ClF3N2O3; molecular weight is 502.9. Its form is a viscous amber oil; its specific gravity is 1.26 at 25°C; log Kow  4.26 or 6.4. It is less soluble (0.001 mg/ l) in water at 25°C, but it is soluble in most organic solvents. It is unstable to light and in alkaline medium. Metabolism The metabolism of fluvalinate has been extensively studied with the 14C-preparation labeled in the acid moiety. When fluvalinate labeled with 14C in the acid moiety was administered to rats at 1 mg/kg, 14C was rapidly excreted into urine (9–19%) and feces (75–88%) within 4 days after administration. The major 14C component (45% of the fecal 14C) was the parent compound. Liver showed relatively higher 14C tissue residue than other tissues, indicating that the 14C tissue residues from fluvalinate are somewhat different from those of other pyrethroids having the

O F

Flumethrin

HOOC CI

C CH

COOH

HOOCCH2NHOC

O F

F

CI Flumethrin acid

glucuronide

O

FPB acid

HOOC

O F 4′ -OH -FPB acid

Figure 76.14  Metabolic pathways of flumethrin in animals.

HOOCCH2NHOC OH

O F

OH

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1650

same ­alcohol moiety. The major metabolic pathways (Figure 76.15) are cleavage of ester linkage and oxidation at the acid and alcohol moieties. Ester hydrolysis leads to formation of anilino acid. The anilino acid is further conjugated with amino acids (glycine, serine, threonine, and valine), bile acids (cholic, taurochoric, and taurochenodeoxycholic), and gly­ cerols (oleoyl- and linoleoylglycerol). In addition, an amide derivative of anilino acid was found. These conjugation reactions of anilino acid with bile acid (cholic acid, taurocholic acid, and taurochenodeoxycholic acid) and with glycerol and monoglycerides have rarely been reported as conjugates with xenobiotics (Quistad et al., 1982, 1983). The major urinary metabolites are anilino acid, its hydroxymethyl derivative, its glycine conjugate, haloaniline, and sulfate conjugate of hydroxyhaloaniline. On the other hand, the major fecal metabolites are anilino acid, its amide derivative, and several conjugates of anilino acid with several endogenous components. An unexpected difference between fluvalinate and other pyrethroids is the minimal amount of hydroxylation at the 4 position of the alcohol moiety. Thus, 4-hydroxyfluvalinate is a very minor fecal metabolite, whereas the amount of 4-hydroxylated parent compound is significant for other pyrethroids – that is, 4-hydroxy-deltamethrin and 4-hydroxycypermethrin (Quistad et al., 1982, 1983). When 14C-fluvalinate labeled in the alcohol moiety was orally administered to rats, the alcohol moiety showed metabolic fates similar to those of pyrethroids having -cyano3-phenoxybenzyl alcohol (Staiger and Quistad, 1984).

CF3

HN Cl

O

COOCH

When 14C-acid-fluvalinate was administered to rhesus monkeys at 1 mg/kg, the 14C was excreted into urine (37%) and feces (55%) within 5 days after administration. The major metabolites found were anilino acid as its hydroxymethyl derivatives and glucuronide. There are several species differences between rats and monkeys: (1) glucuronide conjugation of anilino acid is a major metabolic pathway in rhesus monkeys, whereas little or no glucuronides are detected in rats, (2) conjugation with bile acids is a significant reaction in rats, but it is only a very minor process in rhesus monkeys, and (3) conjugation with glycerol and monoglycerides occurs in rats but is not detected in rhesus monkeys (Quistad and Selim, 1983).

76.17  Imiprothrin Chemical name 2,5-Dioxo-3-(2-prop-2-ynyl)imidazolidin1-ylmethyl (1R)-cis,trans-2,2-dimethyl-3-(2-methylprop-1enyl)cyclopropanecarboxylate. Synonyms Imiprothrin (BSI, E-ISO) is the common name in use. Code designation includes S-41311. The trade name is Pralle. The CAS registry number is 72963-72-5. Physical and chemical properties The empirical formula is C17H22N2O4; molecular weight is 318.4. Its form is a viscous liquid; its specific gravity is 1.1 at 20°C; log Kow  2.9. It is less soluble (93.5 mg/l) in water at 25°C, but it is soluble in most organic solvents. Metabolism When 14C-(1R)-trans- or (1R)-cis-imi­ prothrin labeled in the alcohol moiety was administered

CF3

HN Cl

CN Fluvalinate

Cl HOCH

OH

CONH2 HN

CF3

O

COOCH

O

COOCH

OH

CN

O HN

CF3

CN

amide derivative amino acid conjugate bile acid conjugate glycerols glucuronide

COOH

Cl Anilino acid

refer to fenvalerate metabolism

CH2OH CF3

HN

COOH

CF3

Cl

NH2 Cl

OH CF3

NH2 Cl

Figure 76.15  Metabolic pathways of fluvalinate in animals.

sulfate

Chapter | 76  Pyrethroid Chemistry and Metabolism

1651

orally to rats at 1 or 200 mg/kg, the 14C was rapidly and almost completely eliminated from rats within 7 days after administration (98–103% of the dosed 14C). The urinary 14 C excretion was 83–97%, whereas the fecal excretion was 16% or less. The urinary excretion of 14C for the trans isomer was slightly larger (89–97%) than that for the cis isomer (83–91%). 14C excretion into the expired air was less than 3%. 14C tissue residues on day 7 after administration were generally low in all of the dosed group. There were no marked sex-related differences in the rate of 14C excretion and the 14C tissue residues between either treatment group (Saito et al., 1996). The major metabolic reactions (Figure 76.16) of transand cis-imiprothrin in rats are: (1) cleavage of the ester linkage, (2) cleavage of the imidomethylene linkage, (3) hydroxylation of the imidazolidine ring, (4) dealkylation of the 2-propynyl group, and (5) oxidation at the trans-methyl

group in the isobutenyl side chain (Saito et al., 1995). The major urinary metabolites were 2,4-dioxo-1-(2-propynyl)imidazolidine (PGH), PGH-OH, and hydantoin (HYD) from the trans isomer and these ester-cleaved metabolites and metabolites with intact ester linkage from the cis isomer.

76.18  Kadethrin (RU15525) Chemical name 5-Benzyl-3-furylmethyl (E)-(1R)-cis-2, 2-dimethyl-3-(2-oxothiolan-3-ylidenemethyl) cyclo‑ propanecarboxylate. Synonyms Kadethrin (kadethrine) is the common name in use and is also the trade name. Code designations include RU 15525. The CAS registry number is 58769-20-3. Physical and chemical properties The empirical formula is C23H24O4S; molecular weight is 396.5. Its form O

O

O

O

OH

O

O

OH

HOOC

N

N

COOCH2N O

O OH PGH-OH

OH

w t-acid-cis-PGH-OH

O

O HN

N

HOOC

COOCH2N

N

O w t-acid-cis-PGH

O PGH

O

O

O COOCH2N

COOCH2N

N

HOCH2N

N

N

O

O

O trans-Imiprothrin

cis-Imiprothrin O

O HN

HOOC

NH

COOCH2N

NH

O w t-acid-cis-HYD

O HYD

O

O

O COOCH2N

OH

O

O HN

N

COOCH2N

N

HOCH2N

N

COOCH2N

NH

O

Figure 76.16  Metabolic pathways of imiprothrin in animals.

HOCH2N O

NH

COOCH2N O

NH

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1652

O

O

O S

HOH2C

O

O S

COOCH2

OH

COOCH2

Kadethrin (refer to resmethrin metabolism)

glucuronide

O S

glycine conjugate

COOH O HOOC HS

HOOC

HOOC CH3S

O HOOC CH3S O

COOH

CH3S

COOCH2

HS

COOH

HOOC O

COOH

O HOS

COOH

HOOC O

O

O S COOH

OH

COOCH2

Figure 76.17  Metabolic pathways of kadethrin in animals.

is a yellow-brown viscous oil; log Kow  5.4. It is practically insoluble in water, but it is soluble in most organic solvents. It is unstable to light and in alkaline medium. Metabolism The acid moiety of kadethrin is unique in terms of chemical structure and the alcohol moiety is the same as that of resmethrin. The metabolism of kadethrin has been studied after oral administration of 14C-labeled preparations in the acid or alcohol moiety to rats. The 14C from the acid and alcohol moieties was rapidly and completely excreted into urine and feces (Ohsawa and Casida, 1980). The major metabolic reactions (Figure 76.17) are cleavage of the ester linkage and oxidation at the acid and alcohol moieties. The alcohol moiety after ester hydrolysis is further metabolized as shown in the studies with resmethrin (Miyamoto et al., 1971). The acid moiety initially undergoes hydrolysis of the thiolactone ring, and the resulting mercaptan derivative is oxidized directly to a sulfonic acid or is first methylated before oxidation to a methyl sulfoxide and then a methyl sulfone. Hydroxylation reactions and thiolactone hydrolysis also occur on the intact molecule (Ohsawa and Casida, 1980).

76.19  Metofluthrin Chemical name 2,3,5,6-Tetrafluoro-4-(methoxymethyl)benzyl (EZ)-(1RS)-cis-trans-2,2-dimethyl-3-(1-propenyl)cyclo­ propanecarboxylate. Synonyms Metofluthrin (BSI, ISO) is the common name in use. The trade name is Eminence and SumiOne. Code

designations include S-1264. The CAS registry number is 240494-70-6. Physical and chemical properties The empirical formula is C18H20F4O3; molecular weight is 360.4. Its form is a pale yellow liquid; its specific gravity is 1.21 at 20°C; log Kow  5.0 It is less soluble (0.73 mg/l) in water at 25°C, but it is soluble in most organic solvents. Metabolism Metofluthrin has a new type of chrysanthemic acid (norchrysanthemic acid) in the acid moiety. On single oral administration of metofluthrin labeled with 14 C in the alcohol moiety or in the acid moiety to rats at 1 or 20 mg/kg, almost all of the dosed 14C was eliminated into urine and feces within 7 days after administration, and tissue residues 7 days after administration were very low (Tomigahara, 2005). The major biotransformation reactions (Figure 76.18) of metofluthrin are summarized as follows: (1) oxidation at the methyl group of the propenyl, epoxidation, or reduction at the double bond and oxidation at the dimethyl groups in the acid moiety to form the aldehyde, alcohol, and carboxylic acid derivatives, (2) demethylation at the methoxy group of the alcohol moiety, (3) cleavage of the ester linkage, and (4) conjugation of the resulting metabolites with glucuronic acid, taurine, or mercapturic acid. In rats, metofluthrin induced CYP2B1 through CAR and liver tumor, as is the case with phenobarbital. It is highly likely that liver tumor caused by metofluthrin in rats is similar to that seen for phenobarbital in terms of mode of action, and that this pyrethroid will not have hepatocarcinogenic activity in humans (Deguchi et al., 2009).

Chapter | 76  Pyrethroid Chemistry and Metabolism

O F

OH

F

O

F

F

O

F

O

F

F

O

O F

1653

F

F

O

OH O

F

F

O

F

F

HO

O

F

F

F

F

F

O

O

F

HO

O

O F

O

F

metofluthrin

F

F

O

O

F OH

F

F

F

F

O F

F

F OH

F

F

O

OH O

O F

F

F

HO

F

HO

O

OH

F

mercapturic acid

F

F

F

F

HO

glucuronide

OH O

F

HO

O F

F

OH O F

O

F

O F

HO

F

OH O

F

O F

F

F

F

HO

O

OH

O OH

O

O F

F

F

F

HO

HO

F

F

F

F

F

F

HO

OH

OH

HO

OH

O

O F

F

F

F

HO O

HO O

O

O

O

F

HO

O

OH

HO

OH

O

O F

F

O

OH

mercapturic acid

Figure 76.18  Metabolic pathways of metofluthrin in animals.

76.20  Permethrin Chemical name 3-Phenoxybenzyl (1RS)-cis-trans-3-(2,2dichlorovinyl)-2,2-dimethylcyclopropanecarboxylate. Synonyms Permethrin (BSI, E-ISO, ANSI, ESA, BAN) is the common name in use. There are many trade names, such as Agniban, Ambush, Assithrin, Dragnet, Dragon, Eksmin, Outflank, Pounce, Sanathrin, and Talcord. Code designations include S-3151, PP 557, FMC 33297, NRDC143, WL43479, and LE79-519. The CAS registry number is 52645-53-1. Physical and chemical properties The empirical formula is C21H20Cl2O3; molecular weight is 391.3. Its form is a pale yellow-brown liquid; its specific gravity is 1.29 at 20°C; log Kow  6.1. It is less soluble (0.006 mg/l) in water at 20°C, but it is soluble in most organic solvents. It is more stable in acid medium than in alkaline medium. Metabolism The metabolism of permethrin has been studied in detail in a wide variety of animals in vivo and in vitro (rats, cows, hens, and goats). However, the in vivo metabolism data relating to mammalian toxicology are limited to rats. When the four 14C preparations of (1RS)-trans-, (1R)trans-, (1RS)-cis, or (1R)-cis-permethrin labeled in the alcohol or acid moieties were administered orally to male rats at 1.6–4.8 mg/kg, the compounds were rapidly metabolized and the 14C from the acid and alcohol moiety was almost completely eliminated from the body within a few

days. The 14C from the cis isomer was excreted into the urine and feces almost equally, whereas more than 80% of the dosed 14C from the trans isomer appeared in the urine. The 14C tissue residues were very low, although the fat with the cis isomer showed relatively higher residue levels (Elliott et al., 1976; Gaughan et al., 1977). The major metabolic reactions (Figure 76.19) of both permethrin isomers were oxidation at the trans and cis positions of the gem-dimethyl group of the acid moiety and at the 2 and 4 positions of the alcohol moiety; ester cleavage; and conjugation of the resulting carboxylic acids, alcohols, and phenols with glucuronic acid, glycine, and sulfuric acid. The cis isomer is more stable than the trans isomer, and the cis isomer yielded four fecal ester metabolites that resulted from hydroxylation at the 2 and 4 positions of the phenoxy group, at the trans-methyl group, and at both of the two latter sites. The ester-cleaved metabolites were extensively excreted into the urine, whereas the metabolites retaining ester linkage were found only in the feces. There were no significant differences in metabolism between the (1RS)-isomers and (1R)isomers (Elliott et al., 1976; Gaughan et al., 1977). Rat and mouse liver microsomes hydrolyzed the trans isomers more rapidly than the cis isomers; however, mouse microsomes oxidized the cis isomers more extensively than did rat microsomes. The preferential oxidation sites of the alcohol moiety are the 4 position for rats and the 2, 4, and 6 positions for mice (Shono et al., 1979; Soderlund and Casida, 1977).

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1654

CI CI

HOH C 2

HOH2C

O

COOCH2

OH

4′ -OH-t-per

CI

CI CI

CI

COOH t-CI2CA

CI

CI

sulfate

CI CI CI CI

2

CI

COOCH

2

OH

O

COOCH2

CI

OHC

CH2OH O

CI

t-CH2OH-t-per HOOC

OH

COOH

glucuronide

CI t-CH2OH-c-CI2CA

t-CH2OH-c-per

COOCH

2

O

2′-OH-c-per CH OH

COOH

HOOC

CI

O

OH

PB ald

c-CH2OH-t-CI2CA

c-CI2CA

2

COOCH2

O CI

COOCH2

glucuronide

HOH C

2

CI

COOH

cis-Permethrin

HOH C O

OH

4′-OH, t-OH-c-per

CI

O

O

COOCH2

CI

CI

CI

CH CH 2

CI

4′ -OH-c-per

t-CH2OH-t-per

O

O

COOCH

CI OH

4′ -OH-PB acid

HOH2C

CI

O c-CH2OH-t-CI2CA-lactone

O

PB alc

t-CH2OH-t-CI2CA

CI

OH

4′ -OH-PB alc

HOOC

HOH C CI 2 COOH

CI

O

COOCH2 trans-Permethrin

HOH2C CI

O

CH OH

2

O

CI CI

PB acid

O

2

COOCH

2

c-CH2OH-c-per

O

CI

CI COOH

CI c-CH2OH-c-CI2CA

O

O CI c-CH2OH-c-CI2CA lactone

2′-OH-PB acid HOOCCH2HNOC

O

sulfate PB acid-glycine

Figure 76.19  Metabolic pathways of permethrin in animals.

cis- and trans-Permethrin were good substrates in vitro for rat CYP1A1, 1A2, 2C6, 2C11, and 3A1 and for human CYP1A1, 1A2, 2C9, 2C19, and 3A4 (Scollon et al., 2009). The ester hydrolysis of permethrin isomers is mediated by rat carboxylesterases (ES 3 for cis-permethrin and ES 3,10 for trans-permethrin), human pure carboxylesterases (CE-1 and CE-2), and rabbit carboxylesterase with rapid hydrolysis for the trans isomer. The resulting alcohol (PB alc) is oxidized to PB ald by CYP2C9 and alcohol dehydrogenase, and PB ald is further oxidized to PB acid by CYP1A2, 2B1, 2C6, 2D1, and 3A1 and aldehyde dehydrogenase (Hodgson, 2003; Nakamura et al., 2007; Nishi et al., 2006; Ross et al., 2006).

76.21  Phenothrin (d-Phenothrin) Chemical name 3-Phenoxybenzyl (1RS)-cis-trans-2,2dimethyl-3-(2-methylprop-1-enyl)cyclopropane‑carboxylate. d-Phenothrin is a mixture of two isomers [((1R)-cis) and ((1R)-trans)]. Synonyms Phenothrin (BSI, ISO, BAN) is the common name in use. The trade name is Sumithrin for d-phenothrin. Code designations include S-2539 for phenothrin and S2539 Forte for d-phenothrin. The CAS registry number is 26002-80-2. Physical and chemical properties The empirical formula is C23H26O3; molecular weight is 350.5. Its form is a pale yellow to yellow-brown liquid; its specific gravity is

1.06 at 20°C; log Kow  6.0. It is less soluble (0.01 mg/l) in water at 25°C, but it is soluble in most organic solvents. It is unstable to light and in alkaline medium. Metabolism The metabolism of phenothrin was studied in rats after single oral or dermal application of transand cis-phenothrin labeled with 14C in the alcohol moiety at several doses (Kaneko et al., 1981b; Miyamoto et al., 1974; Suzuki et al., 1976). The major excretion route of 14 C was urine for the trans isomers, whereas it was feces for the cis isomers. The residue levels of 14C were generally very low except in fat, which showed slightly higher residue levels. Dermal absorption rate of 14C was estimated to be 3–17% in rats, depending on the formulation used. In addition, comparative metabolism of the six isomers [(1R)-trans, (1RS)-trans, (1S)-trans, (1R)-cis, (1RS)-cis, and (1S)-cis] of phenothrin was investigated after single oral administration of each 14C-labeled preparation of the six phenothrin isomers to rats and mice at 10 mg/kg (Kaneko et al., 1984c). There were no significant differences in the amount of 14C urinary and fecal excretion between (1R)-trans and (1RS)-trans isomers or between (1R)-cis and (1RS)-cis isomers. On the other hand, the (1S)-trans and (1S)-cis isomers showed slightly larger 14C urinary excretion than the corresponding (1R)- and (1RS)trans and -cis isomers, respectively (Izumi et al., 1984). The major metabolic reactions (Figure 76.20) of the cis isomer in both animals were oxidation at the cis- and transmethyl of the isobutenyl group, at the trans-methyl of the

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1655

O

COOCH2 cis-Phenothrin

trans-Phenothrin

O

HOCH2

Ester metabolites R3

R2

Ester metabolites R1

R1

PB alc COOCH2

R2 CH3 CH2OH COOH CH3 CH2OH CH3 COOH COOH COOH

R2

O R4

R1 1) CH2OH 2) CH3 3) CH3 4) CH2OH 5) CH3 6) COOH 7) CH3 8) CH3 9) CH3

O

COOCH2

R3 CH3 CH3 CH3 CH3 CH3 CH3 CH3 CH2OH CH2OH

R4 H H H OH OH OH OH H OH

O

HOOC

taurine conjugate glucuronide glycine conjugate

PB acid

HOOC

O

HOOC

4’ -OH -PB acid sulfate

sulfate

OH

glucuronide

O

COOCH2

O R3

R1 1) COOH 2) CH2OH

R2 CH3 CH3

R3 H H

OH

2’ -OH -PB acid sulfate

Figure 76.20  Metabolic pathways of phenothrin in animals.

gem-dimethyl group attached to the cyclopropane ring, and at the 4 position of the alcohol moiety. Hydroxylation at the 2 position of the alcohol moiety occurred to a smaller extent. On the other hand, with the trans isomer, the main metabolic reaction was cleavage of ester linkage. The hydroxylation at the 4 position of the alcohol moiety occurred with the trans isomer to the same degree as that with the cis isomer. Moreover, small amounts of the metabolites hydroxylated at the trans-methyl of the iso­ butenyl group and at the 2 position of the alcohol moiety were obtained from the trans isomer. The species differences observed in phenothrin were as follows: (1) 4-OHPB acid–sulfate and PB acid–taurine were characteristic to rats and mice, respectively, as is the case with fenvalerate, deltamethrin, and cypermethrin, and (2) all six phenothrin isomers underwent cleavage of the ester linkage and hydroxylation at the 4 position of the alcohol moiety more rapidly in rats than in mice. These facts may contribute to the higher urinary 14C excretion in rats than in mice (Izumi et al., 1984; Kaneko et al., 1984b; Miyamoto et al., 1974; Suzuki et al., 1976). With respect to metabolic reactions, the (1S)-trans and (1S)-cis isomers underwent ester cleavage to a larger extent than the other chiral isomers in both rats and mice. There were some differences in the extent of oxidation at the 4 position of the alcohol moiety of phenothrin isomers among three trans isomers and among three cis isomers. With the trans isomers, there was no apparent difference in the extent of hydroxylation at the 4 position, whereas the (1S)-cis

isomers underwent hydroxylation to a slightly larger extent compared with the other cis isomers. Based on the results, it may be concluded that the (1R) and (1RS) isomers behaved in a similar metabolic manner, and that the (1S) isomers received cleavage of the ester linkage slightly more readily than the corresponding (1R) and (1RS) isomers, although the (1S) isomers also received the same metabolic reactions as the (1R) and (1RS) isomers. However, purified ­carboxylesterase showed virtually no difference in ester hydrolysis between (1R)-trans and (1S)-trans isomers or between (1R)-cis and (1S)-cis isomers, although it hydrolyzed the (1R)-trans and (1S)-trans isomers more rapidly than the (1R)-cis and (1S)-cis isomers, respectively (Izumi et al., 1984; Suzuki and Miyamoto, 1978).

76.22  Prallethrin Chemical name (S)-2-Methyl-4-oxo-3-prop-2-ynylcyclopent2-enyl (1R)-cis-trans-2,2-dimethyl-3-(2-methylprop-1-enyl)­ cyclopropanecarboxylate. Synonyms Prallethrin (BSI, E-ISO) is the common name in use. The trade name is Etoc. Code designations include S4068 SF. The CAS registry number is 23031-36-9. Physical and chemical properties The empirical formula is C19H24O3; molecular weight is 300.4. Its form is a yellow to yellow-brown liquid; its specific gravity is 1.03 at 20°C; log Kow  4.49. It is less soluble (8 mg/l) in water at 25°C, but it is soluble in most organic solvents.

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1656

Metabolism On single oral or subcutaneous administration of (4S,1R)-trans- or (4S,1R)-cis-prallethrin labeled with 14C in the alcohol moiety to rats at 2 mg/kg, 96–104% of the dosed 14C was eliminated into urine and feces within 7 days after administration. Monitoring of the expired air indicated that less than 0.1% of the dose was excreted as 14 CO2 (Shiba et al., 1988). Urinary excretion of 14C with the trans isomer was larger (60–78%) than that with the cis isomer (17–32%), whereas 14C fecal excretion with the trans isomer was smaller than that with the cis isomer. 14C levels in blood and other tissues reached maximum within 3 h after oral administration and thereafter decreased rapidly. 14C tissue residues were generally very low on day 7 after administration (Shiba et al., 1988). Nineteen metabolites were identified in urine and feces (Shiba et al., 1988). Two new types of S-linked conjugation (sulfonic acid and mercapturic acid types) were additionally identified (Tomigahara et al., 1994c). The major biotransformation reactions (Figure 76.21) of prallethrin are summarized as follows: (1) oxidation at the methyl group of the isobutenyl group in the acid ­ moiety, (2) hydroxylation at the C1 or the C2 position of the propynyl group in the alcohol moiety, (3) cleavage of the ester linkage, and (4) conjugation of the resulting metabolites with glucuronic acid, sulfuric acid, sulfonic acid, or mercapturic acid.

76.23  Pyrethrins Chemical name The six known insecticidally active compounds in pyrethrum are esters of two acids and three alcohols. Specifically, pyrethrin I is the pyrethrolone ester of chrysanthemic acid, pyrethrin II is the pyrethrolone ester of pyrethric acid, cinerin I is the cinerolone ester of chrysanthemic acid, cinerin II is the cinerolone ester of pyrethric acid, jasmolin I is the jasmolone ester of chrysanthemic acid, and jasmolin II is the jasmolone ester of pyrethric acid. Synonyms Pyrethrins (BSI, ISO, JMAF, ESA) is the common name in use. There are several trade names, such as Alfadex, Evergreen, and ExciteR. The CAS registry numbers are 121-21-1 (pyrethrin I), 4466-14-2 (jasmolin I), 25402-06-6 (cinerin I), 121-29-9 (pyrethrin II), 117263-0 (jasmolin II), and 121-20-0 (cinerin II). Physical and chemical properties Empirical formulas and molecular weights: Pyrethrin I Pyrethrin II Cinerin I Cinerine II Jasmolin I Jasmolin II

C21H28O3 C22H28O5 C20H28O3 C21H28O5 C21H30O3 C22H30O5

Their forms may be viscous oils or tan dusts; log Kow  5.9 (pyrethrin I) and 4.3 (pyrethrin II). They are practically

COO

COO O

O

trans-Prallethrin

cis-Prallethrin

mercapturic acid conjugate

sulfonic acid conjugate HO

Ester metabolites R2

R1

328.4 372.4 316.4 360.4 330.4 374.5

Ester metabolites R3

glucuronide

O

R1

R2

COO O R2 CH3

R3 11

2) COOH

CH3

3)

CH3

4) CH2OH

CH3

12

5)

CH3

13

CH3 CH3

HO

HO

R1 1) CH2OH

COO

R2

OH

O

O

O 1)

R1 CH3

R2 CH2OH

R3 11

11

2)

CH3

COOH

11

12

3)

CH3

CH3

12

4)

CH3

CH2OH

12

5)

CH3

CH3

13

O

OH OH glucuronide

glucuronide O

Figure 76.21  Metabolic pathways of prallethrin in animals.

sulfate

Chapter | 76  Pyrethroid Chemistry and Metabolism

1657

insoluble in water, but they are readily soluble in most organic solvents. They are unstable to light and air. Metabolism On single oral administration of pyrethrin I and pyrethrin II labeled with 3H or 14C in the acid or alcohol moiety to rats and mice at 1–5 mg/kg, the 3H and 14C from both compounds were excreted into urine and feces almost equally. The trans-methyl group of the isobutenyl side group of pyrethrin I is readily oxidized to hydroxymethyl derivatives and oxidized further to the corresponding carboxylic acid derivatives through aldehydes. Hydrolyis of the methyl ester of pyrethrin II yields the same acid. This acid is further oxidized at the pentadienyl group of the alcohol moiety, which is probably initiated by epoxidation of the terminal double bond. Hydrolysis of this epoxide resulted in formation of isomeric diols, one of which is conjugated with an unidentified aromatic acid (Elliott et al., 1972). The major metabolic reactions (Figure 76.22) of both compounds were hydrolysis of the methoxycarbonyl group, oxidation at the trans-methyl of the isobutenyl group of the acid moiety, and epoxidation at the pentadienyl side chain to yield initially a 4,5-epoxide from which the two diol derivatives were derived (Casida et al., 1971; Elliott et al., 1972). Pyrethrins (containing six derivatives) and phenobarbital both induced CYP2B and CYP3A in cultured rat hepatocytes and human hepatocytes. The induction of pyrethrins appears to be qualitatively similar to that of phenobarbital. Rat liver and thyroid gland tumor formation caused by both

CH3OOC COO

COO

Pyrethrin l O

Pyrethrin lI

O

HOOC

HOH2C

COO

COO

HO

HOOC

OH

O

HOOC

COO PMA

O

P acid

O

P alc

COO O

conjugate

O

HOOC

OH COO OH PMB

Figure 76.22  Metabolic pathways of pyrethrin in animals.

O

compounds can be associated with the CYP induction, and based on epidemiological data for phenobarbital, these tumors do not occur in humans (Price et al., 2007).

76.24  Resmethrin (bioresmethrin, cismethrin) Chemical name 5-Benzyl-3-furylmethyl (1RS)-cis-trans3-(2-methylprop-1-enyl)-cyclopropanecarboxylate. Synonyms Resmethrin (BSI, E-ISO, ANSI, JMAF, ESA) is the common name in use. Trade names are Chrysron and Synthrin. Code designations include SBP1382, NRDC104, and FMC17370. Bioresmethrin and cismethrin are (1R)-trans and (1R)-cis isomers of resmethrin, respectively. The CAS registry number is 10453-86-8. Physical and chemical properties The empirical formula is C22H26O3; molecular weight is 338.4. Its form is a colorless crystal; its specific gravity is 0.958–0.968 at 20°C; log Kow  5.43. It is less soluble (0.038 mg/l) in water at 25°C, but it is soluble in most organic solvents. It is unstable to air and light. Metabolism The metabolism of resmethrin was studied in rats after single oral administration of (1RS, trans)-, (1R, trans)-, or (1R, cis)-resmethrin labeled with 14C in the alcohol or acid moiety (Miyamoto et al., 1986; Ueda et al., 1975a). In addition, in vitro studies with mouse and rat liver microsomes have also been carried out to elucidate the metabolism of resmethrin in detail (Ueda et al., 1975b). Resmethrin was rapidly metabolized and the 14C from the acid moiety was more rapidly eliminated from the body than that from the alcohol moiety. The major metabolites were 5-benzyl-3-furoic acid (BFCA), 4-OH-BFCA, and -OH-BFCA from the alcohol moiety and hydroxymethyl and dicarboxylic acid derivatives of chrysanthemic acid from the acid moiety. Although the trans-methyl group of the isobutenyl group was predominantly oxidized in bioresmethrin and cismethrin, the cis-methyl group of cismethrin was also oxidized. In addition, epimerization at C3 of the cyclopropane ring was found, leading to isomerized forms of dicarboxylic acid derivatives of chrysanthemic acid. Bioresmethrin was hydrolyzed more rapidly in rat liver microsomes than cismethrin, whereas the plasma did not show this cis-trans specificity (White et al., 1976). The major metabolic reactions (Figure 76.23) of resmethrin are cleavage of the ester linkage; oxidation at the trans- and the cis-methyl group of the isobutenyl group, at the 4 position of the benzene ring, and at the benzylic position of the alcohol moiety; and conjugation of the resulting metabolites with glucuronic acid and sulfuric acid. Resmethrin was a good substrate in vitro for rat CYP1A1, 2C6, 2C11, and 3A1 and for human CYP, 2C8, 2C9, and 2C19 (Scollon et al., 2009). In addition, bioresmethrin was hydrolyzed specifically by human CE1 but not by human CE2 (Ross and Crow, 2007).

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COOH HOOC wt- or wc-acid -c- or t-cA

OHC

COOH

cis trans interconversion initiated by proton abstraction at the C-3 carbon

COOH HOH2C wt- or wc-alcc- or t-CA

COOH O

c- or t-CA

COOCH2

O

Resmethrin

HOH2C

O

BFA

O

glucuronide HOOC

O HOOC

OHC

BFCA

O OH

HOOC

4’-OH -BFCA

O OH

HOOC

glucuronide, sulfate

α-OH -BFCA

O HOOC

OH

O

α-Keto-BFCA

Figure 76.23  Metabolic pathways of resmethrin in animals.

76.25  Tetramethrin (d-teTramethrin) Chemicalname 3,4,5,6-Tetrahydrophthalimidomethyl (1RS)cis-trans-chrysanthemate; d-tetramethrin is composed of the (1R) isomers. Synonyms Tetramethrin (BSI, E-ISO, ANSI) is the common name in use. Trade names are Neo-Pynamin and Duracide for tetamethrin and Neo-Pynamin Forte for d-tetramethrin. Code designations include SP 1103 and FMC 9260. The CAS registry numbers are 7696-12-0; 51348-904 for (1R)-cis isomer and 1166-46-7 for (1R)-trans isomer. Physical and chemical properties (tetramethrin) The empirical formula is C19H25NO4; molecular weight is 331.4. Its form is a colorless crystal; its specific gravity is 1.1 at 20°C; its vapor pressure is 0.944 mPa at 30°C; log Kow  4.6. It is less soluble (1.83 mg/l) in water at 25°C, but it is soluble in most organic solvents. It is unstable in strong acid and alkaline medium. Metabolism In vivo and in vitro rat metabolism of this pyrethroid had been reported several times in the 1960s, 1970s, 1980s, and 1990s. The metabolic fates of the alcohol moiety, the tetrahydrophthalimido group, appeared to

be completed in the 1990s. On the other hand, the metabolic fate of the acid moiety was the same as that of pyrethroids having chrysanthemic acid such as resmethrin (Kaneko et al., 1981c; Ueda et al., 1975a). When 14C-(1RS)-trans- or 14C-(1RS)-cis-tetramethrin labeled in the alcohol moiety was administered orally to rats at 2 or 250 mg/kg, the 14C was almost completely eliminated from rats within 7 days after administration. 14 C recovery in feces and urine was 38–56 and 42–58%, respectively, with the trans isomer and 66–91 and 9–31%, respectively, with the cis isomer. Fourteen metabolites were found in excreta. The main metabolites for both of the isomers were sulfonate derivatives in feces and alcohol derivatives and dicarboxylic acid derivatives derived from the 3,4,5,6-tetrahydrophthalimide moiety in urine. The sulfonic acid conjugates have a sulfonic acid group incorporated into the double bond of the 3,4,5,6-tetrahydrophthalimide moiety. Two of five sulfonic acid conjugates (trans-acid-3-OH-NPY-SA and TPI-SA) were detected in the urine; however, their amounts were smaller than in the feces. In addition, the sulfonic acid conjugates were not detected in the bile or urine of the bile duct-cannulated rat given 14C-alcohol-trans- or cis-tetramethrin. Therefore, it is likely that the sulfonic acid conjugates were produced in the intestinal tract (Tomigahara et al., 1994b). The major metabolic reactions (Figure 76.24) of transand cis-tetramethrin in rats were as follows: (1) cleavage of the ester linkage, (2) cleavage of the imide linkage, (3) hydroxylation of cyclohexene or cyclohexane ring of the 3,4,5,6-tetrahydrophthalimide moiety, (4) oxidation at the methyl group of the isobutenyl moiety of the acid moiety, (5) reduction at the 1,2-double bond of the 3,4,5,6tetrahydrophthalimide moiety, and (6) incorporation of a sulfonic acid group into the 1,2-double bond of the 3,4,5,6tetrahydrophthalimide moiety (Kaneko et al., 1981c; Miyamoto et al., 1968; Tomigahara et al., 1994a,b, 1996). The 14C was rapidly and almost completely excreted into the urine and feces after a single oral dose of each of the 14 C-(1R)-trans, (1RS)-trans, (1R)-cis, and (1RS)-cis isomers labeled in the acid moiety to pregnant rats at 300 mg/kg. The 14 C from both the trans isomers was excreted into the urine to a larger extent than the feces. There were no statistical differences in the 14C excretion between both the trans isomers. The 14C derived from both the cis isomers was almost completely eliminated from rat body for 7 days as with the trans isomers. However, significant differences were observed in the excretion of the 14C into the urine and feces between the (1R)- and (1RS)-cis isomers. There were no appreciable differences in 14C tissue residues and the nature of metabolites between the (1R)- and (1RS)-cis isomers, although there were significant differences in absorption in the gastro­ intestinal tissues between these isomers. From these findings, once entered into the bloodstream, the (1R)- and (1RS)-cis isomers apparently undergo the same metabolic reactions. Overall, there seems to be no significant differences in the

Chapter | 76  Pyrethroid Chemistry and Metabolism

1659

O

O

HOOC

HN

COOCH2N

O

OH

O SO3H trans-Acid-NPY-SA O

HOOC

HOOC

HOOC

O TPI

THPA

THAM

O COOCH2N

COOCH2N

H2NOC

HN O SO3H TPI-SA

O SO3H trans-Acid-NPY-SA HOOC

O

O

O trans-Tetramethrin

HOCH2N

HN O MTI

O

OH

O

O HPI

COOCH2N

HOCH2N O SO3H 3-OH-MTI-SA

O cis-Tetramethrin O

HOOC

OH

COOCH2N O SO3H cis-Acid-3-OH-NPY-SA

O

HOOC

OH

HOOC

HOOC

HOOC

HOOC OH

HN HOOC OH 1-OH-5-oxo-HPA

O

O 3-OH-HPI-1 and 2

TCDA

1-OH-HPA

Figure 76.24  Metabolic pathways of tetramethrin in animals.

metabolic reactions between the trans isomers and between the cis isomers. The nature and amounts of the metabolites detected in the fetus with Neo-Pynamin [a mixture of (1RS)isomers] and Neo-Pynamin Forte [a mixture of (1S)-isomers] were nearly the same, indicating that these two compounds behaved in the same way to give the same placental transfer regardless of their optical differences (Kaneko et al., 1984b).

deltamethrin. After formation of deltamethrin, this compound appears to undergo the same metabolic fates of delta­ methrin. Similarly, tralocythrin was rapidly converted to (1R, S)-cis-cypermethrin by debromination, as is the case with tralomethrin (Cole et al., 1982). The debromination was mediated by tissue thiols such as glutathione (Kaneko et al., 1986b; Ruzo et al., 1981).

76.26  Tralomethtin

Conclusion

Chemical name (S)--Cyano-3-phenoxybenzyl (1R)cis-2,2-dimethyl-3-[(RS)-1,2,2,2-tetrabromoethyl] cyclopropanecarboxylate. Synonyms Tralomethrin (BSI, ANSI, ISO) is the common name in use. Trade names are Saga, Scout, and Tralox. Code designations include RU 25474, NU831, and HAG 107. The CAS registry number is 66841-25-6. Physical and chemical properties The empirical formula is C22H19Br4NO3; molecular weight is 665.0. Its form is an orange to yellow resinous solid; its specific gravity is 1.70 at 20°C; log Kow is approximately 5. It is less soluble (0.080 mg/l) in water at 25°C, but it is soluble in most organic solvents. Metabolism A comparative metabolism study has been carried out on the fate of tralomethrin and deltamethrin in male rats after single oral administration of 14C-tralo‑ methrin and 14C-deltamethrin labeled in the acid moiety, the alcohol moiety, and the CN group at 0.3–0.5 mg/kg (Cole et al., 1982). Tralomethrin was not normally detected in the treated animals or their excreta because it undergoes rapid and substantially complete debromination to form

All the pyrethroid insecticides investigated so far are rapidly metabolized in mammals, and their metabolites are almost completely excreted in the urine and feces within several days of single oral or subcutaneous administration except their cyano moiety. Regarding percutaneous absorption of pyrethroids, humans show much less skin penetration than do rats, indicating that the rat model may lead to overestimation of dermal absorption, although a limited number of experiments have been performed. Tissue residues are generally very low, showing that they are biodegradable and nonbioaccumulative (Miyamoto et al., 1995). This is due partly to rapid metabolism. In addition, this rapid degradation may lead to low mammalian toxicity. The major metabolic reactions of pyrethroids are commonly oxidation of methyl groups and aromatic rings in the molecules, hydrolysis of the ester linkage, and several types of conjugation reactions. With respect to cleavage of the ester linkage, there are significant differences between geometrical trans and cis isomers. The trans isomers of pyrethroids having chrysanthemic acid derivatives in the acid moiety such as phenothrin and permethrin are more rapidly

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hydrolyzed than the corresponding cis isomers, and the cis isomers yield more metabolites retaining intact ester linkage than the trans isomers (Casida and Ruzo, 1980; Miyamoto, 1976, 1981; Ruzo and Casida, 1977). With respect to oxidation, the trans-methyl of the isobutenyl group in chrysanthemates is oxidized preferentially to the cis-methyl group, and the 4 position of the phenoxy ring in the phenoxybenzyl derivatives is more readily oxidized than all other positions. Conjugation reactions of pyrethroids include hydrophilic conjugations with glucuronic acid, sulfuric acid, and amino acids and lipophilic conjugations with cholesterol, bile acid, and triglyceride. The optical isomers of pyrethroids do not show significant differences in metabolism reactions, except that only one isomer of the four optical isomers of fenvalerate produces the cholesterol ester conjugate. Although human data are very limited, there seems to be substantially no remarkable species difference in metabolism reactions between laboratory animals and humans: ester hydrolysis and oxidation reactions are major reactions in laboratory animals and humans. However, some in vitro studies and human studies have shown that there are some species differences in the rate of oxidation reaction and ester hydrolysis. Major metabolic reactions for pyrethroid insecticides are ester hydrolysis and oxidation. These reactions are mediated by carboxylesterases and several CYP isoforms. With respect to ester hydrolysis, rats and humans show a distinct difference in serum carboxylesterases: the esterase activity is high in rat serum, but the activity is lacking in humans. In humans, hCE1 may be a predominant carboxyl­esterase for the ester hydrolysis of pyrethroids, and in rats hydrolase A and B show high esterase activity. With respect to CYP isoforms, rat CYP1A1, 1A2, 2C6, 2C11, 3A1, and 3A2 are involved in oxidation reactions of several pyrethroids; among these, 2C11 and 2C6 are considered to play a predominant role in male rat liver on the basis of their enzyme-specific activity and abundance. In human liver, CYP2C8, 2C9, 2C19, and 3A4 are involved in oxidation metabolism; among these, 2C9 and 3A4 are likely to be major CYP isoforms for oxidation reactions of several pyrethroids based on the specific enzyme activity and the amounts (Scollon et al., 2009). Biomonitoring studies of pyrethroids have been conducted to determine contamination levels of pesticides in nonoperators and operators in the United States, United Kingdom (Llewellyn et al., 1996), Germany (Hardt and Angerer, 2003; Heudorf et al., 2004; Leng et al., 2003), and Japan (Wang et al., 2007). This monitoring was most extensively carried out in the United States, where several urinary metabolites derived from ester hydrolysis of several pyrethroids were analyzed for people with different ages (older than 6 years), gender, and ethnicity by the Centers for Disease Control and Prevention (CDC, 2005). The following metabolites were analyzed: 4-fluoro-3-phenoxybenzoic acid (the alcohol moiety of cyfluthrin), cis-3-(2,2-dichlorovinyl)2,2-dimethylcyclopropane carboxylic acid (the acid moiety

of cyfluthrin, cis-permethrin, and cis-cypermethrin), trans3-(2,2-dichlorovinyl)-2,2-dimethylcyclopropane carboxylic acid (the acid moiety of cyfluthrin, trans-permethrin, and trans-cypermethrin), cis-3-(2,2-dibromovinyl)-2,2-dimethyl­ cyclopropane carboxylic acid (the acid moiety of deltamethrin), and 3-phenoxybenzoic acid (the alcohol moiety of many type II pyrethroids). 3-Phenoxybenzoic acid showed somewhat higher urinary levels than other metabolites. The results showed no significant gender, age, or ethnic differences, with urinary levels being very low, indicating that the levels did not result in adverse health effects (CDC, 2005).

References Anand, S. S., Bruckner, J. V., Haines, W. T., Muralidhara, S., Fisher, J. W., and Padilla, S. (2006). Characterization of deltamethrin metabolism by rat plasma and liver microsomes. Toxicol. Appl. Pharmacol. 212, 156–166. Capt, A., Luzy, A. P., Esdaile, D., and Blanck, O. (2007). Comparison of thee human skin grafted onto nude mouse model with in vivo and in vitro models in the prediction of percutaneous penetration of three lipophilic pesticides. Regul. Toxicol. Pharmacol. 47, 274–287. Casida, J. E., and Ruzo, L. O. (1980). Metabolic chemistry of pyrethroid insecticides. Pestic. Sci. 11, 257–269. Casida, J. E., Kimmel, E. C., Elliott, M., and Janes, N. F. (1971). Oxidative metabolism of pyrethrins in mammals. Nature 230, 326–327. Centers for Disease Control and Prevention (2005). “Third National Report on Human Exposure to Environmental Chemicals”. Centers for Disease Control and Prevention, Atlanta. Chamberlain, K., Matsuo, N., Kaneko, H., and Khambay, B. P. S. (1998). Pyrethroids. In “Chirality in Agrochemicals” (N. Kurihara and J. Miyamoto, eds.), pp. 9–84. Wiley, New York. Class, T. J., Ando, T., and Casida, J. E. (1990). Pyrethroid metabolism: Microsomal oxidation metabolites of (S)-bioallethrin and the six natural pyrethrins. J. Agric. Food Chem. 38, 529–537. Cole, K. M., Ruzo, L. O., Wood, E. J., and Casida, J. E. (1982). Pyrethroid metabolism: Comparative fate in rats of tralomethrin, tralocythrin, deltamethrin and (1 R.aS)-cis-cypermethrin. J. Agric. Food Chem. 30, 631–636. Crawford, M., and Hutson, D. H. (1977). The metabolism of the pyrethroid insecticide ()--cyano-3-phenoxybenzyl 2,2,3,3-tetramethyl­ cyclopropanecarboxylate, WL41706, in the rats. Pestic. Sci. 8, 579–599. Crawford, M. J., Croucher, A., and Hutson, D. H. (1981a). Metabolism of cis- and trans-cypermethrin in rats. Balance and tissue retention study. J. Agric. Food Chem. 29, 130–135. Crawford, M. J., Croucher, A., and Hutson, D. H. (1981b). The metabol­ ism of the pyrethroid insecticide cypermethrin in rats: Excreted metabolites. Pestic. Sci. 12, 399–411. Crow, J. A., Borazjani, A., Potter, P. M., and Ross, M. K. (2007). Hydrolysis of pyrethroids by human and rat tissues: Examination of intestinal, liver and serum carboxylesterases. Toxicol. Appl. Pharmacol. 221, 1–12. Deguchi, Y., Yamada, T., Hirose, Y., Nagahori, H., Kushida, M., Sumida, K., Sukata, T., Tomigahara, Y., Nishioka, K., Uwagawa, S., Kawamura, S., and Okuno, Y. (2009). Mode of action analysis for the synthetic pyrethroid metofluthrin-induced rat liver tumors: Evidence for hepatic CYP2B induction and hepatocyte proliferation. Toxicol. Sci. 108, 69–80.

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Hutson, D. H., and Casida, J. E. (1978). Taurine conjugation in metabol­ ism of 3-phenoxybenzoic acid and the pyrethroid insecticide cypermethrin in mouse. Xenobiotica 8, 565–571. Hutson, D. H., Gaughan, L. C., and Casida, J. E. (1981). Metabolism of the cis- and trans-isomers of cypermethrin in mice. Pestic. Sci. 12, 385–398. International Programme on Chemical Safety (1989). “Allethrins – Allethrin, d-Allethrin, Bioallethrin and S-Bioallethrin.” Environmental Health Criteria 87, World Health Organization, Geneva. International Programme on Chemical Safety (1990a). “Cyhalothrin.” Environmental Health Criteria 99. World Health Organization, Geneva. International Programme on Chemical Safety (1990b). “Deltamethrin.” Environmental Health Criteria 97. World Health Organization, Geneva. Isobe, N., Kaneko, H., Shiba, K., Saito, K., Ito, S., Kakuta, N., Saito, A., Yoshitake, A., and Miyamoto, J. (1990). Metabolism of esfenvalerate in rats and mice and effects of its isomers on metabolic fates of esfenvalerate. J. Pestic. Sci. 15, 159–168. Isobe, N., Suzuki, T., Nishikawa, J., Kaneko, H., Nakatsuka, I., and Yoshitake, A. (1992). Metabolism of empenthrin isomers in rats. J. Pestic. Sci. 17, 27–37. Izumi, T., Kaneko, H., Matsuo, M., and Miyamoto, J. (1984). Comparative metabolism of the six stereoisomers of phenothrin in rats and mice. J. Pestic. Sci. 9, 259–267. Kaneko, H., Ohkawa, H., and Miyamoto, J. (1981a). Comparative metabolism of fenvalerate and the [2S,aS]-isomer in rats and mice. J. Pestic. Sci. 6, 317–326. Kaneko, H., Ohkawa, H., and Miyamoto, J. (1981b). Adsorption and metabolism of dermally applied phenothrin in rats. J. Pestic. Sci. 6, 169–182. Kaneko, H., Ohkawa, H., and Miyamoto, J. (1981c). Metabolism of tetramethrin isomers in rats. J. Pestic. Sci. 6, 425–435. Kaneko, H., Izumi, T., Matsuo, M., and Miyamoto, J. (1984a). Metabolism of fenvalerate in dogs. J. Pestic. Sci. 9, 269–274. Kaneko, H., Izumi, T., Ueda, Y., Matsuo, M., and Miyamoto, J. (1984b). Metabolism and placental transfer of stereoisomers of tetramethrin isomers in pregnant rats. J. Pestic. Sci. 9, 249–258. Kaneko, H., Matsuo, M., and Miyamoto, J. (1984c). Comparative metabol­ ism of stereoisomers of cyphenothrin and phenothrin isomers in rats. J. Pestic. Sci. 9, 237–247. Kaneko, H., Matsuo, M., and Miyamoto, J. (1986a). Differential metabolism of fenvalerate and granuloma formation: I. Identification of a cholesterol ester derived from a specific chiral isomer of fenvalerate. Toxicol. Appl. Pharmacol. 83, 148–156. Kaneko, H., Takamatsu, Y., Kitamura, N., Yoshitake, A., and Miyamoto, J. (1986b). In vivo and in vitro conversion of tralomethrin to delta­ methrin in larvae of tobacco cutworm, Spodoptera litura. J. Pestic. Sci. 11, 533–540. Kaneko, H., Shiba, K., Yoshitake, A., and Miyamoto, J. (1987). Metabolism of fenpropathrin (S-3206) in rats. J. Pestic. Sci. 12, 385–395. Kaneko, H., Takamatsu, Y., Okuno, Y., Abiko, J., Yoshitake, A., and Miyamoto, J. (1988). Substrate specificity for formation of cholesterol ester conjugates from fenvalerate analogues and for granuloma formation. Xenobiotica 18, 11–19. Leng, G., Leng, A., Kuhn, K.-H., Lewalter, J., and Pauluhn, J. (1997). Human dose-excretion studies with the pyrethroid insecticide cyfluthrin: Urinary metabolite profile following inhalation. Xenobiotica 27, 1273–1283. Leng, G., Ranft, U., Sugiri, D., Hadnagy, W., Berger-Preiss, E., and Idel, H. (2003). Pyrethroids used indoors – Biological monitoring of exposure to pyrethroids following an indoor pest control operation. Int. J. Hyg. Environ. Health 206, 85–92.

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Llewellyn, D. M., Brazier, A., Brown, R., Cocker, J., Evans, M. L., Hampton, J., Nutley, B. P., and White, J. (1996). Occupational exposure to permethrin during its use as a public hygiene insecticide. Ann. Occup. Hyg. 40, 499–509. Miyamoto, J. (1976). Degradation, metabolism and toxicity of synthetic pyrethroids. Environ. Health Perspect. 14, 15–28. Miyamoto, J. (1981). The chemistry, metabolism and residue analysis of synthetic pyrethroids. Pure Appl. Chem. 53, 1967–2022. Miyamoto, J., Sato, Y., Yamamoto, K., Endo, M., and Suzuki, S. (1968). Biochemical studies on the mode of action of pyrethroidal insecticides: Part 1. Metabolic fate of phthalthrin in mammals. Agric. Biol. Chem. 32, 628–640. Miyamoto, J., Nishida, T., and Ueda, K. (1971). Metabolic fate of resmethrin, 5-benzyl-3-furylmethyl dl-trans-chrysanthemate in the rat. Pestic. Biochem. Physiol. 1, 293–306. Miyamoto, J., Suzuki, T., and Nakae, C. (1974). Metabolism of phenothrin or 3-phenoxybenzyl d-trans-chrysanthemumate in mammals. Pestic. Biochem. Physiol. 4, 438–450. Miyamoto, J., Kaneko, H., and Takamatsu, Y. (1986). Stereoselective formation of a cholesterol ester conjugate from fenvalerate by mouse microsomal carboxyesterase(s). J. Biochem. Toxicol. 1, 79–94. Miyamoto, J., Kaneko, H., Tsuji, R., and Okuno, Y. (1995). Pyrethroids, nerve poisons: How their risks to human health should be assessed. Toxicol. Lett. 82/83, 933–940. Nakamura, Y., Sugihara, K., Sone, T., Isobe, M., Ohta, S., and Kitamura, S. (2007). The in vitro metabolism of a pyrethroid insecticide, permethrin, and its hydrolysis products in rats. Toxicology 25, 176–184. Nishi, K., Huang, H., Kamita, S. G., Kim, I. H., Morisseau, C., and Hammock, B. D. (2006). Characterization of pyrethroid hydrolysis by the human liver carboxylesterases hCE-1 and hCE-2. Arch. Biochem. Biophys. 445, 115–123. Ohkawa, H., Kaneko, H., Tsuji, H., and Miyamoto, J. (1979). Metabolism of fenvalerate (Sumicidin) in rats. J. Pestic. Sci. 4, 143–155. Ohsawa, K., and Casida, J. E. (1980). Metabolism in rats of the potent knockdown pyrethroid kadethrin. J. Agric. Food Chem. 28, 250–255. Okuno, Y., Seki, T., Ito, S., Kaneko, H., Watanabe, T., Yamada, H., and Miyamoto, J. (1986). Differential metabolism of fenvalerate and granuloma formation: II. Toxicological significance of a lipophilic conjugate from fenvalerate. Toxicol. Appl. Pharmacol. 83, 157–169. Price, R. J., Walters, D. G., Finch, J. M., Gabriel, K. L., Capen, C. C., Osimitz, T. G., and Lake, B. G. (2007). A mode of action for induction of liver tumors by pyrethrins in the rat. Toxicol. Appl. Pharmacol. 218, 186–195. Quistad, G. B., and Selim, S. (1983). Fluvalinate metabolism by rhesus monkeys. J. Agric. Food Chem. 31, 596–599. Quistad, G. B., Staiger, L. E., and Schooley, D. A. (1982). Xenobiotic conjugation: A novel role for bile acids. Nature 296, 462–464. Quistad, G. B., Staiger, L. E., Jamieson, G. C., and Schooley, D. A. (1983). Fluvalinate metabolism by rats. J. Agric. Food Chem. 31, 589–596. Roberts, T., and Hutson, D. (1999). “Metabolic Pathways of Agrochemicals: Part 2. Insecticides and Fungicides.” Royal Society of Chemistry Information Services, Cambridge, UK. Ross, K., and Crow, J. A. (2007). Human carboxylesterases and their role in xenobiotics and endobiotic metabolism. J. Biochem Mol. Toxicol. 21, 187–196. Ross, M. K., Borazjani, A., Edwards, C. C., and Potter, P. M. (2006). Hydrolytic metabolism of pyrethroids by human and other mammalian carboxylesterases. Biochem. Pharmacol. 71, 657–669. Ruzo, L. O., and Casida, J. E. (1977). Metabolism and toxicology of pyrethroids with dihalovinyl substituents. Environ. Health Perspect. 21, 285–292.

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Ruzo, L. O., Unai, T., and Casida, J. E. (1978). Decamethrin metabolism in rats. J. Agric. Food Chem. 26, 918–924. Ruzo, L. O., Engel, J. L., and Casida, J. E. (1979). Decamethrin metabolites from oxidative, hydrolytic and conjugative reactions in mice. J. Agric. Food Chem. 27, 725–731. Ruzo, L. O., Gaughan, L. C., and Casida, J. E. (1981). Metabolism and degradation of the pyrethroids tralomethrin and tralocythrin in insects. Pestic. Biochem. Physiol. 15, 137–142. Saito, K., Kaneko, H., Tomigahara, Y., Nakatsuka, I., and Yamada, H. (1995). Metabolism of imiprothrin isomers in rats: Biotransformation and excretion. J. Pestic. Sci. 20, 529–540. Saito, K., Kaneko, H., Tomigahara, Y., Nakatsuka, I., and Yamada, H. (1996). Metabolism of imiprothrin isomers in rats: Absorption and distribution. J. Pestic. Sci. 21, 49–55. Scollon, E. J., Starr, J. M., Godin, S. J., DeVito, M. J., and Hughes, M. F. (2009). In vitro metabolism of pyrethroid pesticides by rat and human hepatic microsomes and cytochrome P450 isoforms. Drug Metab. Dispos. 37, 221–228. Seguchi, K., Asaka, S., Katoh, Y., and Yamaguchi, I. (1991). Metabolism of cycloprothrin in rats. J. Pestic. Sci. 16, 591–598. Shiba, K., Kakuta, N., Kaneko, H., Nakatsuka, I., Yoshitake, A., Yamada, H., and Miyamoto, J. (1988). Metabolism of the pyrethroid insecticide S-4068F in rats. J. Pestic. Sci. 13, 557–569. Shiba, K., Kaneko, H., Kakuta, N., Yoshitake, A., and Miyamoto, J. (1990). Placental transfer of esfenvalerate and fenvalerate in pregnant rats. J. Pestic. Sci. 15, 169–174. Shono, T., Ohsawa, K., and Casida, J. E. (1979). Metabolism of transand cis-permethrin, trans and cis-cypermethrin and deltamethrin by microsomal enzymes. J. Agric. Food Chem. 27, 316–325. Soderlund, D. M., and Casida, J. E. (1977). Effects of pyrethroid structure on rate of hydrolysis and oxidation by mouse liver microsomal enzymes. Pestic. Biochem. Physiol. 7, 391–401. Staiger, L. E., and Quistad, G. B. (1984). Fluvalinate metabolism in rats. J. Agric. Food Chem. 32, 1130–1133. Suzuki, T., and Miyamoto, J. (1978). Purification and properties of pyrethroid carboxyesterase in rat liver microsome. Pestic. Biochem. Physiol. 8, 186–198. Suzuki, T., Ohno, N., and Miyamoto, J. (1976). New metabolites of ()cis fenothrin, 3-phenoxybenzyl ()-cis chrysanthemumate, in rats. J. Pestic. Sci. 1, 151–152. Tomigahara, Y., Mori, M., Shiba, K., Isobe, N., Kaneko, H., Nakatsuka, I., and Yamada, H. (1994a). Metabolism of tetramethrin isomers in rat: I. Identification of a sulphonic acid type of conjugate and reduced metabolites. Xenobiotica 24, 473–484. Tomigahara, Y., Mori, M., Shiba, K., Isobe, N., Kaneko, H., Nakatsuka, I., and Yamada, H. (1994b). Metabolism of tetramethrin isomers in rat: II. Identification and quantitation of metabolites. Xenobiotica 24, 1205–1214. Tomigahara, Y., Shiba, K., Isobe, N., Kaneko, H., Nakatsuka, I., and Yamada, H. (1994c). Identification of two new types of S-linked conjugates of Etoc in rat. Xenobiotica 24, 839–852. Tomigahara, Y., Onogi, M., Miki, M., Yanagi, K., Shiba, K., Kaneko, H., Nakatsuka, I., and Yamada, H. (1996). Metabolism of tetra­ methrin isomers in rat: III. Stereochemistry of reduced metabolites. Xenobiotica 26, 201–210. Tomigahara, Y., Nagahori, H., Matsui, M., Tarui, H., Isobe, N., Kawamura, S., Kaneko, H., and Mikami, N. (2005). Metabolism of metofluthrin in rats. Abstract of the annual meeting of the Pesticide Science Society of Japan. Tomlin, C. D. S. (2006). “A World Compendium, The Pesticide Manual,” 14th ed. British Crop Protection Council, Berkshire, UK.

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Ueda, K., Gaughan, L. C., and Casida, J. E. (1975a). Metabolism of ()trans-and ()-cis-resmethrin in rats. J. Agric. Food Chem. 23, 106–115. Ueda, K., Gaughan, L. C., and Casida, J. E. (1975b). Metabolism of four resmethrin isomers by liver microsomes. Pestic. Biochem. Physiol. 5, 280–294. Ujihara, K., Mori, T., Iwasaki, T., Sugano, M., Shono, Y., and Matsuo, N. (2004). Metofluthrin: A potent new synthetic pyrethroid with high vapour activity against mosquitoes. Biosci. Biotechnol. Biochem. 68, 170–174. Wang, D., Kamijima, M., Imai, R., Suzuki, T., Kameda, Y., Asai, K., Okamura, A., Naito, H., Ueyama, J., Saito, I., Nakajima, T., Goto, M., Shibata, E., Kondo, T., Takagi, K., Takagi, K., and Wakusawa, S.

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(2007). Biological monitoring of pyrethroid exposure of pest control workers in Japan. J. Occup. Health 49, 509–514. White, I. N. H., Verschoyie, R. D., Moradian, M. H., and Barnes, J. M. (1976). The relationship between brain levels of cismethrin and bioresmethrin in female rats and neurotoxic effects. Pestic. Biochem. Physiol. 6, 491–500. Woollen, B. H., Marsh, J. R., Laird, W. J. D., and Lesser, J. E. (1992). The metabolism of cypermethrin in man: Differences in urinary metabolic profiles following oral and dermal administration. Xenobiotica 22, 983–991.

Chapter 77

Toxicology and Mode of Action of Pyrethroid Insecticides David M. Soderlund Cornell University, Geneva, New York

77.1  Introduction Pyrethroids are a class of synthetic insecticides that have been designed and optimized based on the structures of the pyrethrins (Figure 77.1), the six insecticidal constituents of the natural insecticide pyrethrum (Elliott, 1995). Pyrethrum is arguably the most effective natural insecticide, but its use is limited by its instability in light and air, which limits its effectiveness in crop protection and other insect control contexts in which residual activity is essential. The development of pyrethroids involved an iterative process of structural modification and biological evaluation in an effort to identify compounds with increased photostability that retained the potent and rapid insecticidal activity and relatively low acute mammalian toxicity of pyrethrum. The registration in the late 1970s of the first pyrethroids with sufficient environmental stability for agricultural O

applications led to the rapid expansion of the use of these compounds as replacements for older organochlorine and organophosphorus insecticides. Although originally developed and registered for agricultural markets, pyrethroids are also used to control vectors of human disease and are the most common class of insecticides in household insecticide products available to the general public. In 2002, pyrethroid use represented approximately 18% of the U.S. dollar value of the world insecticide market, ranking second only to organophosphorus compounds among insecticide classes (Pickett, 2004). This chapter provides a review of the toxicological properties of pyrethroid insecticides in mammals, with emphasis on the neurotoxic properties that are characteristic of pyrethroids as a class. It is intended to complement the previous chapter on pyrethroid chemistry and metabolism (see Chapter 76). O

H

H O

O O

O

O

Pyrethrin I O

O

Pyrethrin II O

H O

O O

O O O

Cinerin I O

H

Cinerin II O

H O

H O

O

Jasomolin I

O

O O

Jasmolin II

Figure 77.1  Structures of the six natural pyrethrins. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

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77.2  Chemistry and insecticidal action 77.2.1  Development of Synthetic Pyrethroids Synthetic pyrethroids were conceived and developed by modifying the structures of the pyrethrins (Figure 77.1) to increase photostability while retaining the potent and rapid insecticidal activity and relatively low acute mammalian toxicity of pyrethrum (Elliott, 1995). Most synthetic pyrethroids were discovered by the sequential replacement of structural elements of the pyrethrins with novel structural moieties that were selected to conserve the molecular shape and physical properties of the template structure. Because the pyrethrins are esters of a cyclopropanecarboxylic acid and a cyclopentenolone alcohol (Figure 77.2), synthetic modifications typically held one of these major domains of the molecule constant while introducing new structural

Acid moiety

Alcohol moiety

O

Pyrethrin I

H O O

O

Resmethrin

O O

O

Permethrin

O

O Cl Cl Br

O

H

CN O

Deltamethrin

Br

features in the other. Although the early stages of this process employed the natural pyrethrins as templates, subsequent stages employed newly discovered synthetic pyrethroids with desirable insecticidal activity, stability, and other properties as templates for the further design of new compounds. Important advances in the development of synthetic pyrethroids are illustrated by the compounds shown in Figure 77.2. Replacement of the cyclopentenolone ring of the pyrethrin and allethrin alcohols with an alternative unsaturated heterocyclic moiety gave resmethrin (Figure 77.2), which not only exhibited increased photostability but also was substantially more potent as an insecticide and lower in acute mammalian toxicity than pyrethrin I. The combination of these desirable properties in a single molecule provided a strong impetus to search for new compounds with greater activity and photostability. Further developments led to permethrin (Figure 77.2), the first synthetic pyrethroid with sufficient photostability for agricultural use. Compared to resmethrin, this compound contains structural replacements in both the alcohol moiety (3-phenoxybenzyl for 5-benzyl-3-furylmethyl) and the acid moiety (chlorines for methyl groups) that confer enhanced photostability without loss of insecticidal activity. Inclusion of an -cyano substituent in the 3-phenoxybenzyl alcohol moiety, as in deltamethrin (Figure 77.2), produced compounds with much greater insecticidal potency than permethrin but with similar photostability. Synthetic pyrethroids related in structure to permethrin and deltamethrin constitute the largest chemical subfamily of pyrethroids in current use. The structural diversity of synthetic pyrethroids was further enhanced by the discovery that the 2,2dimethylcyclo­propanecarboxylic acid moiety of the pyrethrins and most synthetic compounds could be replaced by an -isopropylphenylacetic acid moiety. This new series of compounds led to the discovery of the commercial insecticide fenvalerate (Figure 77.2). More radical changes based on this series included replacement of the central ester bond to yield compounds such as etofenprox (Figure 77.2) that retained the overall configuration of the molecule and exhibited pyrethroid-like insecticidal activity.

O

77.2.2  Structure–Activity Relationships O

Fenvalerate

Cl

Etofenprox

CN O

O

C2H5O

Figure 77.2  Structural evolution of synthetic pyrethroids.

O

O

Pyrethroids as diverse structurally as pyrethrin I and etofenprox conform to a single, operationally defined structure– activity relationship for insecticidal activity that is based on the physical properties, shape, and three-dimensional configuration of the entire molecule (Elliott et al., 1974a). It is evident that there is no specific substructure, reactive entity, or molecular moiety that can be identified as the toxophore that confers pyrethroid-like insecticidal activity. Instead, such activity apparently results from the appropriate fit of the entire molecule at the site of action. The significance of overall molecular shape in the action of pyrethroids is reinforced by the stringent stereospecificity of

Chapter | 77  Toxicology and Mode of Action of Pyrethroid Insecticides

insecticidal action, illustrated in Figure 77.3 for the isomers of resmethrin. Two chiral centers at carbon-1 and carbon-3 of the chrysanthemic acid moiety of pyrethrin I produce two pairs of diastereomers, which are designated trans and cis based on the orientation of the C-1 and C-3 substituents in relation to the plane of the cyclopropane ring (Figure 77.3). The acid moieties of the natural pyrethrins are exclusively in the 1R,trans configuration. When esters were prepared from the four resolved chrysanthemic acid isomers, those with the R configuration at cyclopropane C-1 were insecticidal, whereas the enantiomeric 1S compounds, though physically identical, were without insecticidal activity (Elliott et al., 1974a). This profound stereospecificity extends to compounds such as fenvalerate (see Figure 77.2), in which the 2S configuration of the non-cyclopropane acid moiety is structurally congruent with 1R cyclopropanecarboxylates and gives insecticidal esters, whereas the corresponding 2R esters are inactive. Stereoisomerism is a less common feature of pyrethroid alcohol moieties. Nevertheless, when a chiral center is present in the alcohol moiety at the carbon bearing the hydroxyl group, as in the pyrethrins (Figure 77.1) and esters of -cyano-3-phenoxybenzyl alcohol (e.g., deltamethrin; see Figure 77.2), only one epimer has high insecticidal activity even when esterified to an acid moiety that contains the appropriate stereochemical configuration for high insecticidal activity (Elliott et al., 1978).

77.2.3  Mechanism of Insecticidal Activity Pyrethroid insecticides were optimized using assays of insecticidal activity that were biased toward the detection of the rapid paralysis typical of pyrethrum, an effect presumed to result from action at one or more neuronal targets. Knowledge of the mechanism of insecticidal action of pyrethroids is therefore a useful point of departure for understanding pyrethroid neurotoxicity in other taxa.

Insecticidal

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Pyrethroids alter the normal function of insect nerves by modifying the kinetics of voltage-sensitive sodium channels, which mediate the transient increase in the sodium permeability of the nerve membrane that underlies the rising phase of the nerve action potential. The action of pyrethroids on insect and other invertebrate sodium channels, and the correlation of these effects with insecticidal activity, has been reviewed extensively (Bloomquist, 1993a; Narahashi, 1992, 1996; Sattelle and Yamamoto, 1988; Soderlund and Bloomquist, 1989). A more detailed summary of the actions of pyrethroids on sodium channels is presented in Section 77.7 of this chapter. Confirmation that voltage-sensitive sodium channels are the principal sites of insecticidal action of pyrethroids has emerged from studies of the molecular genetics of mechanisms of pyrethroid resistance that involve reduced nerve sensitivity (reviewed in Soderlund, 2005; Soderlund and Knipple, 2003). In the house fly, the kdr (knockdown resistance) and super-kdr (enhanced knockdown resistance) traits confer resistance to all pyrethroids by reducing the sensitivity of the fly nervous system to these compounds. Genetic linkage analyses mapped these traits close to the principal voltage-sensitive sodium channel gene of the house fly (designated Vssc1). Moreover, kdr-like resistance in other insect species has been mapped to sodium channel genes that are orthologous to Vssc1. DNA sequence analyses of Vssc1 coding sequences from susceptible and resistant house fly strains identified a single mutation common to all resistant strains and a second mutation found only in the highly resistant super-kdr strains. Insertion of these mutations into a susceptible Vssc1 cDNA, followed by functional analysis of the susceptible and specifically mutated sodium channels, showed that these mutations were sufficient to account for the resistance caused by the kdr and super-kdr traits. Therefore, the sodium channels encoded by the Vssc1 gene of the house fly and by orthologous genes of other insect species must be the principal sites of insecticidal action of pyrethroids.

Noninsecticidal

O

O O

O O

1R,trans

O 1S,trans

O

O O

O O

1R,cis

Figure 77.3  Insecticidal and noninsecticidal isomers of resmethrin.

O 1S,cis

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77.3  Acute neurotoxic actions in mammals

the expression of acute toxicity (Gray and Soderlund, 1985). These observations suggest that the acute toxicity of pyrethroids is strongly influenced by toxicokinetic factors. Further evidence for this is found in comparative studies of the oral toxicity of pyrethroids using different dosing vehicles. Administration in aqueous suspensions gives much lower acute oral toxicities than administration in vegetable oils (Soderlund et al., 2002), presumably because the low aqueous solubility of pyrethroids limits their bioavailability when administered in aqueous suspensions. Pyrethroids are rapidly metabolized in mammals in vivo by hydrolysis of the central ester bond and oxidative attack at several sites to give complex arrays of primary and secondary metabolites (see Chapter 76). The use of compounds that interfere with pyrethroid biotransformation as synergists in toxicity studies confirms the significance of biotransformation as a limiting factor in the expression of pyrethroid toxicity. Pretreatment with the esterase inhibitor S,S,Stributyl phosphorotrithioate (DEF) or the cytochrome P450 monooxygenase inhibitor piperonyl butoxide increases the intraperitoneal toxicity of several pyrethroids to mice (Gray and Soderlund, 1985). Differences between pyrethroids in the degree of synergism observed in these studies suggest that the relative importance of hydrolytic and oxidative attack in limiting toxicity varies from pyrethroid to pyrethroid and from isomer to isomer of any given pyrethroid.

77.3.1  Acute Toxicity Pyrethroids are commonly regarded as relatively safe insecticides. However, the acute toxicity of pyrethroids to mammals varies widely with structure. Table 77.1 summarizes the oral toxicity to rats of several pyrethroids that are currently registered for use in the United States. With few exceptions, these compounds have acute oral LD50 values following administration in vegetable oils between 50 and 500 mg/kg and are therefore considered to be moderately toxic (EPA Category II). Pyrethroid intoxication following single doses is transient. Depending on the route of administration, signs of toxicity are observed from minutes to a few hours after dosing. Animals that survive acute intoxication at near-lethal doses recover and appear normal 1– 14 days after treatment (Soderlund et al., 2002). The toxicity of an individual pyrethroid varies widely depending on the route of administration and the vehicle used for dosing. Pyrethroids have very low acute toxicities following dermal exposure due to their limited absorption through the skin (Clark, 1995). In contrast, administration by routes that favor delivery to the central nervous system (e.g., intravenous or intracerebral injection) greatly enhances

Table 77.1  Acute Toxicity to Rats, Structural Classification, and Signs of Intoxication for Selected Pyrethroid Insecticides Oral LD50 (mg/kg)a

Compound

Structural class

Intoxication syndromeb

Males

Females

S-Bioallethrin

  370

  320

Type I

T

Bifenthrin

   70

   54

Type I

–c

Cyfluthrin

  155

  160

Type II

–

-Cyhalothrin

   79

   56

Type II

–

Cypermethrin

  297

  372

Type II

CS

Deltamethrin

   95

   87

Type II

CS

Esfenvalerate

   87

   87

Type II

CS

Fenpropathrin

   71

   67

Type II

T/CS

Permethrin

1200

1200

Type I

T

Pyrethrins

  710

  320

Type I

T

Resmethrin

1695

1640

Type I

T

Tefluthrin

   22

   35

Type I

–

a

Data from Soderlund et al. (2002). Intoxication syndromes as described by Verschoye and Aldridge (1980); T/CS indicates elements of both syndromes present. c Compound not included in classification study (Verschoyle and Aldridge, 1980). b

Chapter | 77  Toxicology and Mode of Action of Pyrethroid Insecticides

Acute intoxication by pyrethroids is correlated directly with levels of the parent pyrethroid in the central nervous system (CNS). The CNS concentrations of pyrethroids obtained following intracerebral injection of severely toxic or lethal doses correspond closely to those found after the peripheral administration of equally toxic doses (Gray and Soderlund, 1985). This relationship was illustrated clearly in studies of deltamethrin intoxication in mice by different routes of administration and following varied synergist pretreatment regimes (Ruzo et al., 1979). Whereas the observed LD50 values for deltamethrin varied over a 30-fold range, the levels of deltamethrin in the CNS at death varied over only a 2.6-fold range. These results imply that targets in the CNS mediate the acute toxicity of pyrethroids irrespective of the route of administration.

77.3.2  Structure–Toxicity Relationships Determination of the acute toxicity of the resolved optical isomers of several pyrethroids following intravenous administration to rats (Verschoyle and Aldridge, 1980) and intracerebral administration to mice (Lawrence and Casida, 1982) permitted the elucidation of structure–activity relationships for acute mammalian toxicity. The neurotoxicity of pyrethroids to mammals depends on the stereochemical configuration at cyclopropane C-1 (see Figure 77.3) or the homologous position in compounds lacking the cyclopropanecarboxylate moiety. Only esters of 1R cyclopropanecarboxylates and isosteric 2S isomers of non-cyclopropane acids are neurotoxic, whereas the corresponding 1S cyclopropanecarboxylates and their sterically equivalent 2R acyclic analogues are without measurable toxicity even when administered at high doses directly to the CNS (Gray and Soderlund, 1985). This relationship parallels the stereospecificity of insecticidal action of pyrethroids (see Section 77.2.3). In mammals, the absolute configuration at cyclopropane C-3 of cyclopropanecarboxylate esters of primary alcohols (e.g., resmethrin, permethrin) also strongly influences toxicity. Typically, pyrethroids from this group of compounds having the 1R,cis configuration (e.g., 1R,cis isomers of resmethrin and permethrin) are both insecticidal and toxic to mammals, whereas the corresponding isomers having the 1R,trans configuration, though similar in insecticidal potency, lack measurable acute toxicity to mammals. Initially, the low toxicity of these 1R,trans compounds to mammals was ascribed to their rapid hydrolytic detoxication by liver esterases, but intracerebral dosing experiments (Lawrence and Casida, 1982) demonstrated that these compounds had very low intrinsic toxicities even when the impact of biotransformation was removed. In mammals, as in insects, the presence of an -cyano substituent in S configuration in the 3-phenoxybenzyl alcohol moiety also greatly enhances acute neurotoxicity. In the case of mammalian toxicity, this effect is highly stereospecific: the -R epimers of compounds that retain the appropriate configurations for high toxicity in the acid moiety have no

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demonstrable toxicity when injected directly into the brain (Ghiasuddin and Soderlund, 1985). The -cyano substituent also indirectly alters structure–toxicity relationships for the acid moiety. The most dramatic effects are seen with the 1R,trans cyclopropanecarboxylates of 3-phenoxybenzyl alcohol (e.g., [1R,trans]-permethrin), which exhibit extremely low toxicity to mammals; addition of an -cyano substituent in the S configuration to these esters produces compounds (e.g., [1R,trans,S]-cypermethrin) with significant neurotoxicity to rodents (Gray and Soderlund, 1985).

77.3.3  Two Syndromes of Pyrethroid Intoxication Studies published in the early 1970s identified two distinct syndromes associated with the acute toxicity of pyrethroids to rats. Verschoyle and Barnes (1972) provided the first systematic description of the signs of pyrethroid intoxication in rats following oral and intravenous dosing. These authors noted the same syndrome of intoxication for pyrethrins, bioallethrin, and resmethrin by either route of administration. This syndrome included hypersensitivity and aggression followed by stimulus-induced bouts of general tremor, convulsive twitching, coma, and death. The principal difference observed between oral and intravenous dosing was the speed of onset of intoxication. The publication of the discovery of deltamethrin (Elliott et al., 1974b), the first pyrethroid containing the -cyano-3-phenoxybenzyl moiety, was accompanied by a brief report describing the acute toxicity of deltamethrin to rats (Barnes and Verschoyle, 1974). This report noted that the signs of deltamethrin intoxication following either oral or intravenous administration, which involved salivation without lacrimation followed by jerking leg movements and progressive writhing convulsions (choreoathetosis), were distinctly different than those reported by these authors for other pyrethroids (Verschoyle and Barnes, 1972). A landmark study (Verschoyle and Aldridge, 1980) described both the acute toxicity and the signs of intoxication of 36 pyrethroids following intravenous administration, thereby establishing a taxonomy of pyrethroid intoxication in mammals that persists to the present. Of the 18 esters of various primary alcohols examined in this study, 15 compounds produced signs of intoxication corresponding to those first described for pyrethrins and pyrethroids (Verschoyle and Barnes, 1972), which was designated the T (tremor) syndrome, whereas the remaining three compounds did not produce any detectable signs of intoxication at the highest doses tested. Of the 17 esters of -cyano-3-phenoxybenzyl alcohol examined, 12 produced signs of intoxication like those first described for deltamethrin (Barnes and Verschoyle, 1974), which was designated the CS (choreoathetosis with salivation) syndrome, whereas four produced the T syndrome of intoxication. One -cyano-3-phenoxybenzyl ester and one compound in which the -cyano group

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was replaced by an -ethynyl group were found to produce elements of both syndromes (tremor with salivation). This classification of the signs of pyrethroid intoxication into two principal syndromes was confirmed in studies of the intracerebral toxicity of 29 pyrethroids to mice (Lawrence and Casida, 1982). An alternative nomenclature (Type I and Type II) has been proposed for subgroups of pyrethroids based not only on the syndromes of intoxication produced in mammals (Lawrence and Casida, 1982) but also on their chemical structures, their signs of poisoning in insects, and their actions on insect nerve preparations (Gammon et al., 1981). Type II pyrethroids contain the -cyano-3-phenoxybenzyl moiety (e.g., deltamethrin and fenvalerate; see Figure 77.2), whereas Type I compounds comprise a wide structural variety of compounds lacking the -cyano-3-phenoxybenzyl group (e.g., pyrethrin I, resmethrin and permethrin; see Figure 77.2). The Type I/Type II nomenclature has been widely adopted in the literature and is often used in a manner parallel to the T/CS nomenclature, so that Type I compounds are generally considered to produce the T syndrome of intoxication and Type II compounds are considered to produce the CS syndrome (Lawrence and Casida, 1982). Although the initial descriptions of the T and CS syndromes suggested that the signs of intoxication were independent of the route of administration (Verschoyle and Aldridge, 1980; Verschoyle and Barnes, 1972), it is not clear from available data that this generalization extends to all pyrethroids or to all routes of administration. For example, comparison of the results of regulatory neurotoxicity studies for nine pyrethroids (Soderlund et al., 2002) showed that the behavioral signs of intoxication following oral exposure were not well correlated with the signs characteristic of the T and CS syndromes following intravenous dosing (see Section 77.4.4).

77.3.4  Neurochemical Consequences of Pyrethroid Intoxication Pyrethroid-dependent neurotransmitter release from presynaptic nerve terminals in the brain was first documented in rats treated with deltamethrin (Aldridge et al., 1978). Deltamethrin treatment resulted in a significant decrease in acetylcholine levels in whole brain and, most significantly, in the cerebellum. In contrast, cismethrin produced no significant reduction in acetylcholine levels (Aldridge et al., 1978). Other effects of pyrethroids in the cerebellum include deltamethrin- and cypermethrin-induced increases in cyclic guanosine monophosphate levels (Aldridge et al., 1978; Brodie and Aldridge, 1982; Brodie and Opacka, 1987; Lock and Berry, 1981). Pyrethroids also increased the levels of some amino acid neurotransmitters and metabolites of monoamine neurotransmitters in the brain. At doses that produced tremor in rats, permethrin increased the levels of

Hayes’ Handbook of Pesticide Toxicology

aspartate in the brain stem and striatum, the levels of glutamate in the brain stem, the levels of serotonin metabolites in the hypothalamus, brain stem, and hippocampus, and the levels of dopamine metabolites in the striatum (Hudson et al., 1986). Similarly, allethrin, cypermethrin, fenvalerate, and permethrin increased the levels of dopamine metabolites in the striatum (Doherty et al., 1986b). Selective effects of pyrethroids on the cerebellum were also found in assays of brain glucose levels during intoxication. Rat brain glucose levels and brain blood flow were elevated in all brain areas except the cerebellum during writhing following a lethal dose of deltamethrin. In contrast, glucose levels in the cerebellum were depleted, indicating a high level of glucose utilization in this region (Cremer et al., 1980, 1983). Cypermethrin produced similar increases in cerebellar glucose levels during writhing but resulted in elevated cerebellar lactate concentrations during both writhing and salivation (Lock and Berry, 1981). Brain blood flow was also increased significantly by cismethrin, but, except for the cerebellum, this increase was eliminated by blocking cortical vasodilation by atropine (Cremer et al., 1983).

77.3.5  Age-Related Differences in Pyrethroid Sensitivity Information on differences in pyrethroid sensitivity between immature and adult animals is limited. Neonatal rats were six- to 17-fold more sensitive than adults to the pyrethroids permethrin, cypermethrin, and deltamethrin (Cantalamessa, 1993; Sheets et al., 1994). The use of synergists to block metabolic detoxication (Cantalamessa, 1993) and the measurement of concentrations in whole brain at equitoxic doses (Sheets et al., 1994) suggested that the greater sensitivity of young animals was due to age-related differences in pharmacokinetics rather than intrinsically greater sensi­tivity of the neonatal CNS. A subsequent study failed to identify age-related differences in the sensitivity of rats to cismethrin (the 1R,cis isomer of resmethrin) or permethrin at either high (lethal) or low (behaviorally active) doses (Sheets, 2000). The same study also found differential sensitivity to deltamethrin and cypermethrin only at high doses. It is possible that the age-related differences observed at high doses in these studies results are due to the reduced ability of the incompletely developed detoxication enzyme systems of the neonatal liver to metabolize a large bolus of insecticide.

77.3.6  Reports of Neurotoxic Effects in Humans Despite three decades of extensive worldwide use, there are surprisingly few documented instances of acute pyrethroid intoxication in humans. The principal source of information

Chapter | 77  Toxicology and Mode of Action of Pyrethroid Insecticides

on human intoxication by pyrethroids is a comprehensive review of 573 cases of acute pyrethroid poisoning that were reported in the Chinese medical literature during the period 1983–1988 (He et al., 1989). These cases encompassed both occupational exposure as the result of mishandling during agricultural uses (229 cases) and accidental exposure, usually by ingestion of formulated insecticide products (344 cases). All but seven of these cases involved three pyrethroids: deltamethrin (325 cases), fenvalerate (196 cases), and cypermethrin (45 cases). The most common signs of systemic pyrethroid poisoning included dizziness, headache, nausea, anorexia, and fatigue, whereas more serious cases exhibited coarse muscle fasciculations, disturbance of consciousness, coma, and convulsive attacks. Seven deaths were reported, of which one was due to atropine intoxication following misdiagnosis of organophosphate insecticide poisoning. All patients received symptomatic and supportive therapies, and most recovered within 6 days of exposure, although recovery of patients experiencing convulsions required up to 55 days. Other reports of isolated cases of pyrethroid exposure and intoxication (Box and Lee, 1996; Gotoh et al., 1998; Lessenger, 1992) are in substantial agreement with the findings reported in the Chinese literature (He et al., 1989). Although lethal effects of pyrethroids are rare, data collected by poison control centers show a significant incidence of pyrethroid exposure. During a recent 4-year period (2001–2004) the Toxic Exposure Surveillance System of the American Association of Poison Control Centers received 55,610 reports of exposure to pyrethroid insecticides, of which nearly one-third (16,946; 30.5%) were exposures to children younger than 5 years of age (Litovitz et al., 2002; Watson et al., 2003, 2004, 2005). Only a small proportion of these exposures (2433; 4.4%) were characterized as producing a moderate or major effect, and only three exposures were fatal. In the two fatalities for which details were reported, pyrethroid exposure was accompanied by exposure to an organic solvent (xylene or ethylene glycol), presumably as a component of the formulated insecticide, making it unclear whether death was due to insecticide or solvent intoxication (Bradberry et al., 2005).

77.4  Behavioral neurotoxicity Several studies have studied the effects of sublethal doses on parameters of pyrethroid behavior. A recent review (Wolansky and Harrill, 2008) provides a comprehensive analysis of the behavioral neurotoxicity of pyrethroids. The following sections summarize data from the most comprehensive of these studies.

77.4.1  Effects on Motor Activity Crofton and Reiter (Crofton and Reiter, 1984, 1988a,b) compared the effects of nine commercial or experimental

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pyrethroids on motor activity in rats. Compounds were administered orally in corn oil and motor activity was assessed in a figure-eight maze at the time of peak effect (1.5–2 h after dosing as defined in these studies). All nine compounds caused a dose-dependent decrease in motor activity, and no differences in this response were noted between compounds identified as producing the T syndrome of intoxication and those producing the CS syndrome. A subsequent comparative study of 11 pyrethroids (Wolansky et al., 2006) confirmed the universal reduction in motor activity by all pyrethroids regardless of symptomatic class and established relative potencies based on doses required to produce a 30% reduction in motor activity. The 11 pyrethroi­ds examined in this study differed by ~240-fold in their relative potencies for the reduction of motor activity. The potency of deltamethrin in these assays varied more than 200-fold depending on the vehicle and route employed (Crofton et al., 1995), a result that is in good agreement with the effects of route and vehicle on the lethal effects of pyrethroids (see Section 77.3.1). A review of 18 studies of pyrethroid-induced effects on motor activity (Wolansky and Harrill, 2008) showed that the results of most studies were consistent with the effects of pyrethroids on motor activity summarized previously.

77.4.2  Effects on Acoustic Startle Response Crofton and Reiter (Crofton and Reiter, 1984, 1988a,b) also assessed the effects of nine pyrethroids on the acoustic startle response of rats. Compounds were administered orally in corn oil and the latency to onset and peak amplitude of startle responses to standard acoustical stimuli were assessed 1.5–2 h after dosing. In contrast to effects on motor activity, effects on acoustic startle response appeared to discriminate between pyrethroids that produce the T and CS syndromes. Compounds identified as producing the T syndrome of intoxication increased the amplitude of the startle response without affecting its latency, whereas compounds classified as producing the CS syndrome of intoxication produced varied effects. In the latter group, cypermethrin and deltamethrin decreased the amplitude and increased the latency of the startle response, whereas fenvalerate increased the amplitude, did not affect latency, and also, unlike any other compound tested, increased sensitization to background acoustical stimuli. Among the three compounds in this study not formally classified according to the T/CS taxonomy of intoxication syndromes, cyfluthrin and flucythrinate decreased the amplitude and increased the latency of acoustic startle responses in a manner similar to cypermethrin and deltamethrin, whereas fluvalinate had no effect on any measure of the acoustic startle response. In another study (Hijzen and Slangen, 1988), cypermethrin, fenfluthrin, and permethrin increased the amplitude of the acoustic startle response in rats, whereas deltamethrin decreased the amplitude of this response. The effects of permethrin and

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deltamethrin in this study are generally consistent with the findings of Crofton and Reiter (Crofton and Reiter, 1984, 1988a,b), but the increase in startle amplitude caused by cypermethrin is an unexpected and contradictory finding.

77.4.3  Effects on Conditioned Behavior Intraperitoneal administration of allethrin, deltamethrin, fenvalerate, or permethrin to rats caused a dose-dependent reduction in the frequency of a previously learned behavior (i.e., bar pressing reinforced by food) (Bloom et al., 1983; Stein et al., 1987). Oral administration of cypermethrin, deltamethrin, or permethrin to rats caused similar reductions in food-reinforced learned behavior (Glowa, 1986; Peele and Crofton, 1987). Daily oral administration of deltamethrin to rats for 15 days also reduced learning and memory measured in a Y maze using a negatively reinforced visual discrimination response (Husain et al., 1996).

Hayes’ Handbook of Pesticide Toxicology

Despite these limitations, it was possible to reach some conclusions from these results (Soderlund et al., 2002). For example, deltamethrin and cyfluthrin produced signs of intoxication that conform generally to the CS syndrome, whereas three compounds (permethrin, S-bioallethrin, pyrethrum) produced a mixture of signs associated with the T and CS syndromes, and three (bifenthrin, cyhalothrin, cypermethrin) produced signs of intoxication that did not clearly correspond to either syndrome. It is noteworthy that none of the compounds identified by Verschoyle and Barnes (Verschoyle and Aldridge, 1980) as producing the T syndrome (permethrin, S-bioallethrin, pyrethrum) caused signs of intoxication in regulatory neurotoxicology studies that were exclusively correlated with the T syndrome. Thus, the T/CS classification system of intoxication syndromes does not adequately describe the various combinations of effects seen following oral administration FOB assessments employed in regulatory neurotoxicity studies.

77.4.4  Results of Regulatory Neurotoxicity Studies

77.5  Neurotoxic effects following dermal exposure

Acute and subchronic adult neurotoxicity screening studies, conducted in accordance with the regulatory guidelines provided by the U.S. EPA, provide a unique opportunity to compare the actions of different pyrethroids determined under standardized experimental conditions. As a screen for neurotoxicity, these studies rely on neurobehavioral tests that include clinical observations, a procedurally standardized functional observational battery (FOB), measurements of motor activity in an automated device, and terminal microscopic examination of neural tissues. Soderlund et al. (2002) reviewed and summarized regulatory neurotoxicity studies for nine registered pyrethroids. These studies were performed in several different laboratories, using similar, but not identical, procedures. The FOB used in each of these studies is based on the procedure described by Moser (Moser, 1989). Comparison of the results obtained in these nine individual studies showed that each pyrethroid produced a distinct combination of neurobehavioral effects at the highest dose tested. Moreover, comparisons of the findings for any given pyrethroid between acute and subchronic studies and of the results for two isomeric compositions of cypermethrin that were examined independently illustrate the consistency and reproducibility of the signs of intoxication derived from the FOB. However, a detailed assessment of how closely these pyrethroids compare with each other or with either classical syndrome of intoxication (i.e., T or CS) is limited by seve­ral factors, including: a lack of concurrence between the signs reported in intravenous toxicity studies (Verschoyle and Aldridge, 1980) with observations used to characterize the effects in the regulatory studies by the oral route; differences in dose selection; and the absence of data for several registered pyrethroids.

The experience of workers involved in the handling of technical or formulated pyrethroids, during either manufacture or use, provides insight into the effects of pyrethroids on humans following cutaneous exposure. Two review articles (Clark, 1995; Vijverberg and van den Bercken, 1990) summarize occupational exposure data from published clinical reports and unpublished reports from industrial sources. The most frequently reported symptom in worker exposure studies was paresthesia, which was characterized by numbness, itching, burning, or tingling of the skin following dermal exposure to a pyrethroid. These sensations generally occurred in the absence of erythema, edema, vesiculation, or other signs of overt skin irritation and were usually limited to the directly exposed areas of the skin. Pyrethroid-induced paresthesia was transient and reversible within hours after exposure, but in some instances it lasted for up to 48 h. No clinical signs of acute pyrethroid intoxication were observed in any of these cases. Reports of occupational exposure are supported and amplified by the results of studies with human volunteers and experimental animals. In human volunteers, the effects of four pyrethroids were evaluated by application to the earlobe (Flannigan and Tucker, 1985). In this study, permethrin caused the least pronounced and flucythrinate the most pronounced paresthesia, whereas cypermethrin and fenvalerate were intermediate and approximately equal in effect. In each case, paresthesia developed within 30 min of exposure, peaked by 8 h, and dissipated by 24 to 32 h after exposure. In studies with guinea pigs, the onset of abnormal sensation caused by six pyrethroids (cypermethrin, deltamethrin, esfenvalerate, fenvalerate, flucythrinate, and permethrin)

Chapter | 77  Toxicology and Mode of Action of Pyrethroid Insecticides

was judged by an increase in scratching, licking, or biting behavior at the site of dermal application (Cagen et al., 1984). The onset of symptoms usually occurred within 1 h after application. However, permethrin elicited a qualitatively lower overall response than the other five compounds tested, which were approximately equal in effect under these experimental conditions. The latency and duration of the behavioral response were affected by the formulation employed for any single pyrethroid, but the magnitude of the response was independent of formulation. The results of worker exposure data coupled with the results of controlled experiments with human volunteers and animals show that paresthesia appears to be an exclusively local effect. It occurs only at the site of dermal exposure, is not correlated with the appearance of a rash or other signs of classical skin irritation, and is not associated with any signs of systemic intoxication. Based on these observations and the proposed origins of paresthesia in the sensory nervous system (Rizzo et al., 1996), pyrethroid-induced paresthesia has been postulated to be a direct excitatory effect of pyrethroids on small sensory nerve fibers in the skin rather than a response due to classical skin irritation.

77.6  Developmental neurotoxicity The reported exposure of children and pregnant women to low levels of pyrethroids (see Section 77.3.6) has raised concerns about the sublethal effects of pyrethroids on the developing nervous system. This issue has been addressed experimentally in studies of the effect of postnatal exposure in rodents on brain neurochemistry and behavior. The earliest studies of this type examined the effects in mice of exposure to the pyrethroids bioallethrin and deltamethrin during the second and third postnatal weeks, a period of rapid brain development, on open-field behavioral parameters and muscarinic acetylcholine receptor density in the brain at the age of 4 months (Ahlbom et al., 1994; Eriksson and Fredriksson, 1991; Eriksson et al., 1993; Eriksson and Nordberg, 1990). These studies consistently identified increases in motor activity without habituation and a downregulation of muscarinic receptor expression in the brain in 4-month-old mice that had been exposed to bioallethrin or deltamethrin during postnatal days 10–16. However, attempts to replicate these results in another laboratory failed to reproduce the effects of either bioallethrin or deltamethrin on behavior or muscarinic receptor expression (Muhammad et al., 2003). A recent review of 22 developmental neurotoxicity studies in rodents (Shafer et al., 2005) did not identify a consistent pattern of pyrethroid-dependent effects on the developing rodent CNS. Whereas studies completed to date do not rule out developmental neurotoxic effects of pyrethroids, additional well-designed and well-controlled experiments are required to identify a causal connection between early sublethal exposure and persistent effects in mature animals.

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77.7  Actions on voltage-gated sodium channels The compelling evidence for effects on voltage-gated sodium channels as the mechanism of insecticidal activity of pyrethroids and the strong conservation of sodium channel structure, function, and pharmacology across animal taxa implicate the voltage-gated sodium channels of the CNS as important sites of action in mammals. Studies over several decades using a variety of techniques have defined the actions of pyrethroids on both invertebrate and mammalian sodium channels. The following sections summarize the most important findings from the perspective of understanding pyrethroid neurotoxicity in mammals.

77.7.1  Electrophysiologtical Studies The actions of pyrethroid insecticides on sodium channels in invertebrate and vertebrate nerve preparations have been widely documented over the past three decades and are extensively and critically summarized in numerous reviews (Bloomquist, 1993a, 1996; Clark, 1995; Narahashi, 1992, 1996; Soderlund, 1995). Intracellular recordings of action potentials demonstrate that pyrethroids produce one of two distinct types of effect on nerve excitability depending on the structure of the pyrethroid employed (Lund and Narahashi, 1983). Type I compounds produce long trains of action potentials (burst discharges) following a single stimulus with little or no effect on resting potential. In contrast, Type II compounds do not induce repetitive firing but instead cause a use-dependent block of the action potential coupled with depolarization of the resting potential. Some compounds exhibit intermediate properties, causing burst discharges of declining amplitude that lead eventually to nerve depolarization and block. Effects of pyrethroids on sodium channel function that underlie these effects on nerve excitability have been elucidated using voltage clamp and patch clamp techniques. Under voltage clamp, all neuroactive pyrethroids retard the kinetics of sodium channel activation, inactivation, and deactivation (reviewed in Bloomquist, 1993a; Narahashi, 1996). This prolongation is evident during a depolarization step as the induction of a slowly developing and persistent sodium current and after repolarization as a slowly decaying sodium tail current. Type I compounds produce tail currents that decay relatively rapidly, whereas Type II compounds produce extremely persistent tail currents. These effects on macroscopic currents are in good agreement with more limited studies at the single-channel level. In patch clamp experiments, tetramethrin, a Type I compound, increase the mean open time of individual sodium channels approximately 10-fold. In contrast, the Type II compounds deltamethrin and fenvalerate increase the mean open times of individual sodium channels in patch clamp studies up to 200-fold and produced channels that remained open at the end of

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depolarizing pulses. Deltamethrin also delayed channel opening in response to a depolarizing pulse. These findings have been interpreted as evidence that pyrethroids stabilize multiple sodium channel states and slow the transitions between states.

77.7.2  Differential Sensitivity of Sodium Channel Isoforms Most of the available information on the actions of pyrethroids on mammalian sodium channels was obtained using neuronal tissue preparations, which are now known to express multiple sodium channel isoforms. As a result, the action of pyrethroids has not been correlated with the expression of identified sodium channel isoforms in these tissues. However, a limited number of physiological studies suggest that sodium channel isoforms expressed in various mammalian tissues exhibit differential sensitivity to pyrethroids. The clearest evidence of differential sensitivity to pyrethroids between sodium channel isoforms is found in the responses of the TTX-sensitive and TTX-resistant sodium channel populations in dorsal root ganglion (DRG) neurons to pyrethroids. The TTX-resistant current in these cells is much more sensitive than the TTX-sensitive current to allethrin (Ginsburg and Narahashi, 1993), tetramethrin (Song et al., 1996; Tatebayashi and Narahashi, 1994), and deltamethrin (Tabarean and Narahashi, 1998). Pore-forming  subunits of voltage-gated sodium channels in mammals are encoded by a family of nine genes (Catterall et al., 2005a). These channels, designated Nav1.1–Nav1.9, are differentially distributed in excitable cells and exhibit unique functional and pharmacological properties (Figure 77.4). The Nav1.1, Nav1.2, Nav1.3, and Nav1.6 isoforms are expressed in the CNS (Goldin, 2001) and represent potential targets for the systemic neurotoxic

actions of pyrethroids. In the most cases, sodium channel  subunits are coexpressed with either one or two auxiliary  subunits that modulate channel gating and kinetics and regulate channel expression (Meadows and Isom, 2005). In mammals, there are four sodium channel  subunits (designated 1–4). Typically, a single neuron expresses multiple sodium channel  and  subunits and therefore may contain several distinct heteromultimeric complexes. Overlapping patterns of sodium channel  subunit expression in the CNS (Felts et al., 1997; Whitaker et al., 2001) limit the utility of native neuronal preparations to identify isoform-dependent differences in pharmacology. This limitation can be overcome by the heterologous expression of cloned individual sodium channel isoforms in the unfertilized oocytes of the frog Xenopus laevis or in transfected mammalian cell lines. The available information on the sensitivity of individual mammalian sodium channel isoforms to pyrethroids is based primarily on expression studies using the Xenopus oocyte system. Rat Nav1.2 sodium channels, which are abundantly expressed in the adult CNS, exhibit very low sensitivity to modification by deltamethrin and other pyrethroids (Smith and Soderlund, 1998; Vais et al., 2000a). In contrast, rat Nav1.8 channels, which are resistant to block by tetrodotoxin (TTX) and restricted in distribution to the peripheral nervous system, are sensitive to modification by a wide structural variety of pyrethroids (Choi and Soderlund, 2006; Smith and Soderlund, 2001; Soderlund and Lee, 2001). Recent studies show that the rat Nav1.3 isoform is much more sensitive to modification by pyrethroids of the Type II structural class than the rat Nav1.2 isoform (Meacham et al., 2008; Tan and Soderlund, 2009b). The sensitivity of the Nav1.3 isoform is of particular interest because it is preferentially expressed in the embryonic and early postnatal rodent CNS (Felts et al., 1997) and may therefore be an important target for developmental neurotoxic effects attributed to pyrethroids (Shafer et al., 2005). Preliminary studies suggest that the Nav1.6 isoform, the

Name

Expression pattern

Nav 1.9 (NaN)

PNS

Nav 1.8 (SNS/PN3)

PNS

Nav 1.5 (H1/SkM2)

cardiacmuscle

Nav 1.4 (µ1/SkM1)

skeletal muscle

Nav 1.1 (brain I) Nav 1.2 (brain II/IIa)

CNS, PNS CNS

Nav 1.3 (brain III)

CNS(early)

Nav 1.6 (NaCh6/PN4)

CNS, PNS

Nav 1.7 (PN1)

PNS,CNS

Figure 77.4  Dendrogram depicting the evolutionary relationships among identified rat sodium channel  subunit isoforms calculated from published amino acid sequences, older synonyms for each isoform, and predominant sites or developmental stages of expression (based on Catterall et al., 2005a).

Chapter | 77  Toxicology and Mode of Action of Pyrethroid Insecticides

most abundantly expressed isoform in the adult brain, is also substantially more sensitive to pyrethroids than the Nav1.2 isoform (Tan et al., 2008). Sodium channel  subunit genes are highly conserved, so that orthologous subunits in rats and humans are 95% identical at the level of amino acid sequence (Goldin, 2001). However, this degree of sequence conservation still results in 50–100 amino acid sequence differences between orthologous channel proteins and could conceivably produce species differences in functional and pharmacological properties. The oocyte expression system offers the opportunity to compare directly the properties and pyrethroid sensitivities of orthologous rat and human sodium channels. The first study of this type compared the sensitivities of orthologous rat and human Nav1.3 sodium channels, which differ at only 54 of 1951 amino acid residues (Tan and Soderlund, 2009b). Surprisingly, human Nav1.3 channels were significantly less sensitive to modification by the pyrethroid insecticide tefluthrin than rat Nav1.3 channels. These results have important implications for understanding the value and limitations of toxicological studies in rats as the basis for assessing risk to humans.

77.7.3  State-Dependent Actions   of Pyrethroids The majority of studies of pyrethroid actions on sodium channel gating under voltage clamp conditions have been performed by equilibrating channels with pyrethroids at hyperpolarized membrane potentials and assessing the effects of pyrethroids upon depolarization. This approach is biased toward the detection of closed-state modification, and most of these studies have not considered the contribution of other channel states to pyrethroid action. However, there is evidence from studies with voltage-clamped squid and crayfish axons that modification by some pyrethroids is enhanced by repetitive depolarization, implying that channel opening increases the affinity of the pyrethroid binding site for these compounds (Brown and Narahashi, 1992; de Weille et al., 1988; Salgado and Narahashi, 1993). The discovery that modification by cypermethrin and deltamethrin of cloned insect sodium channels expressed in Xenopus laevis oocytes is absolutely dependent on repeated depolar­ ization (Smith et al., 1998; Vais et al., 2000b) has led to the conclusion that these compounds bind preferentially to open sodium channels and raised awareness of the significance of state-dependent modification. Several recent studies of the action of pyrethroids on individual rat sodium channel isoforms expressed in oocytes have evaluated the impact of repeated channel activation on modification by pyrethroids. The most extensive study examined 11 structurally diverse pyrethroids as modifiers of Nav1.8 channels in oocytes (Choi and Soderlund, 2006). All 11 compounds produced clearly detectable closed-state modification. In addition, deltamethrin and three structurally related

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compounds produced statistically significant use-dependent enhancement of modification. These four compounds also produced modified channels that exhibited the slowest kinetics of activation and deactivation. It is possible, therefore, that the apparent use dependence found in this study is due to the accumulation of persistently open channels upon repeated depolarization rather than to enhanced binding of these compounds to the open state of the Nav1.8 channel. Information on the state-dependent modification of channel isoforms expressed in the CNS is limited primarily to two compounds, deltamethrin and tefluthrin. Studies with the rat Nav1.2, Nav1.3, and Nav1.6 isoforms expressed in oocytes have identified distinctive and consistent effects of tefluthrin and deltamethrin that are independent of the relative sensitivities of these isoforms to pyrethroids (Tan and Soderlund, 2009a,b). With all three channel isoforms, tefluthrin produced clearly detectable resting modification that was enhanced two- to fourfold by repeated depolarization. With the Nav1.2 and Nav1.6 isoforms, modification by deltamethrin was only detectable following repeated depolarization. Finally, studies of the action of S-bioallethrin on Nav1.6 sodium channels found no enhancement of resting modification upon repeated depolarization (Tan and Soderlund, 2009a). These results suggest that the importance of usedependent modification of brain sodium channel isoforms varies from compound to compound but may be relatively consistent across isoforms and independent of the relative overall sensitivity of these isoforms to pyrethroids.

77.7.4  The Pyrethroid Receptor on Sodium Channels Biochemical studies of the actions of pyrethroids on sodium channels in the brain have elucidated the relationship between the pyrethroid receptor site and other identified binding domains of the sodium channel. Pyrethroids alone do not affect radiosodium uptake into brain synaptosomes, but they allosterically enhance sodium uptake that is stimulated by veratridine or batrachotoxin (Bloomquist and Soderlund, 1988; Ghiasuddin and Soderlund, 1985). Pyrethroids also allosterically enhance the binding of [3H]batrachotoxinin A-20--benzoate (BTX-B) to brain sodium channels and synergistically enhance the allosteric effects of -scorpion toxins and brevetoxins on BTX-B binding (Brown et al., 1988; Lombet et al., 1988; Trainer et al., 1993). The results of these studies implied the existence of a pyrethroid binding site on sodium channels that is distinct from the sites labeled by other radioligands and allosterically coupled to the veratridine/ batrachotoxin, -scorpion toxin, and brevetoxin binding sites. Binding studies using [3H]RU58487, a pyrethroid photoaffinity radioligand, demonstrated high-affinity saturable binding to brain sodium channels that exhibits the allosteric coupling to other binding domains predicted by previous studies (Trainer et al., 1997). Moreover, [3H]RU58487 specifically labeled the

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sodium channel  subunit protein as demonstrated by immunoprecipitation. These studies provide direct evidence for a pyrethroid binding site that is intrinsic to the  subunit protein and is allosterically coupled to sites for other sodium channel ligands. Point mutations in insect sodium channel genes that are associated with pyrethroid resistance provide additional and unique insight into the sodium channel domains that are important for pyrethroid action and may therefore be involved in pyrethroid binding. Comparisons of partial or complete coding sequences of the para-orthologous sodium channel genes from pyrethroid-sensitive and pyrethroid-resistant strains of 16 arthropod species identified 23 unique amino acid substitutions at a total of 19 sequence positions that are associated with pyrethroid resistance (Soderlund, 2005; Soderlund and Knipple, 2003). Studies employing site-directed mutagenesis and expression in oocytes have confirmed that 11 point mutations (at seven sequence positions) modify the sensitivity of insect sodium channels to pyrethroids (Figure 77.5), whereas the remaining 13 mutations remain uncharacterized with respect to function. The majority of resistance-associated mutations are found in or adjacent to the intracellular linkers between transmembrane helices S4 and S5 (designated L4-5) or within transmembrane helices S5 and S6. Current models of the structure of sodium channels and related voltage-gated potassium channels place the L4-5, S5, and S6 segments of each homology domain in close proximity to each other at or near the inner mouth of the channel, with the S6 helices involved in activation gating of the channel (Lipkind and Fozzard, 2000; Shrivastava et al., 2004; Zhao et al., 2004). The existence of multiple

I

pyrethroid resistance mutations in these regions suggests that the pyrethroid receptor may be located close to the inner mouth of the channel. In particular, the majority of confirmed sites of resistance mutations, including all three sites at which multiple amino acid substitutions have been identified, are located in the L4-5, S5, and S6 segments of homology domain II (see Figure 77.5). This observation has led to models postulating a pyrethroid receptor formed primarily by residues in these segments of homology domain II, with some contribution from the S6 helices and associated regions of homology domains I and III (Lee and Soderlund, 2001; Tan et al., 2005; Vais et al., 2003). Recently, information on resistance mutations has been incorporated into a new structural model of the pyrethroid receptor of the house fly Vssc1 sodium channel (O’Reilly et al., 2006). The open configuration of the pore region of the Vssc1 channel, formed by the L4-5, S5, and S6 segments of homology domains I–IV, was modeled based on the crystal structures of homologous voltage-sensitive potassium channels. Computer-aided docking of pyrethroid ligands and the predicted impact of resistance mutations in these domains on ligand docking yielded a putative pyrethroid receptor formed from four residues in the L4-5 and S5 segments of homology domain II (see Figure 77.5): M918, L925, T929, and L932. Interestingly, the amino acid residues in mammalian sodium channels that align with the putative elements of the pyrethroid receptor are absolutely conserved across all nine mammalian sodium channel isoforms. Therefore, this model, which is specific to insect sodium channels, provides little insight into the molecular basis for the differential sensitivity of mammalian sodium channel isoforms to pyrethroids.

L1014F L1014H L1014S

II V410M

III

IV F1538I

outside S5 S1 S2 S3 S4 S5

S6

S1 S2 S3 S4

S1 S2 S3 S4 S5

S6

H N 3

S6

S1 S2 S3 S4 S5

S6

CO2

inside +

C785R

M918T M918V

L932F L932C T929I L925I

Figure 77.5  Sodium channel gene mutations associated with knockdown resistance shown in relation to the extended transmembrane structure of a voltage-sensitive sodium channel  subunit. Circles denote mutated amino acids known to reduce the pyrethroid sensitivity of insect channels expressed in Xenopus oocytes (Soderlund and Knipple, 2003). Mutations are labeled according to the sequence positions of the corresponding residues of the most abundant splice variant of the house fly Vssc1 sodium channel. Filled circles identify mutations considered to form the pyrethroid receptor of the Vssc1 sodium channel (O’Reilly et al., 2006).

Chapter | 77  Toxicology and Mode of Action of Pyrethroid Insecticides

77.7.5  Correlation of Sodium Channel Effects with Toxicity The effects of pyrethroids on voltage-gated sodium channels are correlated both qualitatively and quantitatively with acute intoxication in rodents (Bloomquist, 1993a). At a qualitative level, most Type I compounds, which induce burst discharges in neurons following a single stimulus and rapidly decaying sodium tail currents under voltage clamp conditions, produce the T syndrome of intoxication in rodents. Most Type II compounds, which cause use-dependent nerve block resulting from a degradation of the resting membrane potential and long-lived sodium tail currents under voltage clamp, produce the CS syndrome of intoxication. A small number of compounds from both the Type I and Type II structural classes produce intermediate effects on neuronal excitability and tail current decay and also produce signs of intoxication that include elements of both the T and CS syndromes. The duration of modified sodium currents produced by various pyrethroids is also strongly correlated with acute toxicity. Thus, Bloomquist (1993a) noted a robust correlation between the duration of pyrethroid-induced sodium tail currents measured in frog neurons and both the acute toxicity and signs of intoxication produced by the same compounds in rodents. In light of the evidence connecting effects on sodium channels with toxicity, it is surprising to some observers that the potency of pyrethroids in sodium channel assays, typically measured as the concentration required to produce a half-maximal effect, is relatively modest. Biochemical and electrophysiological studies of pyrethroid action require relatively high concentrations of insecticide to produce populations of pyrethroid-modified channels large enough to be detected by the methods of assay. However, the concentrations required to disrupt normal action potential generation in neurons are orders of magnitude lower because only a small fraction of the available population of channels must be modified in order to disrupt electrical signaling. This relationship was demonstrated directly by Song and Narahashi (1996) in a study of the action of tetramethrin on sodium channels in rat Purkinje neurons. In this study, micromolar concentrations of tetramethrin were required to produce sufficient channel modification (5–25%) to be detected readily under voltage clamp conditions, whereas disruption of action potential generation in the same cells was achieved at 100 nM tetramethrin, a concentration calculated to modify fewer than 1% of sodium channels based on voltage clamp data. Similarly, concentrations of deltamethrin in the micromolar range are required to produce detectable modification of various rat sodium channel isoforms expressed in oocytes under voltage clamp conditions (Choi and Soderlund, 2006; Meacham et al., 2008; Tan and Soderlund, 2009a,b; Vais et al., 2000a), but deltamethrin concentrations in the nanomolar range are sufficient to disrupt normal electrical activity in primary cultures of mouse

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frontal cortex and spinal cord neurons (Shafer et al., 2008). Thus, studies performed in some in vitro systems tend to underestimate significantly the potency of pyrethroids as disruptors of neuronal signaling in intact nervous systems.

77.8  Actions of Pyrethroids on other Neuronal targets The studies summarized in the previous section strongly implicate voltage-gated sodium channels as sites of action for the neurotoxic effects of pyrethroids in mammals as well as in insects. However, the actions of pyrethroids on other neuronal targets have also been investigated, and some of the effects identified in these studies may be relevant to the neurotoxic actions of at least some pyrethroids. In particular, actions on alternative or secondary targets have been invoked to account for the production of the CS intoxication syndrome by most pyrethroids of the Type II structural class. Soderlund et al. (2002) provided a critical analysis of the toxicological significance of pyrethroid action on alternative neuronal targets. The sections that follow consider the actions of pyrethroids on three putative alternative targets (voltage-gated calcium channels, voltage-gated chloride channels, and GABAA receptors) and assess the possible toxicological relevance of these effects, taking into consideration information published since the previous critical review (Soderlund et al., 2002).

77.8.1  Actions of Pyrethroids on VoltageGated Calcium Channels Voltage-gated calcium channels are ubiquitous in excitable membranes and modulate a wide range of cellular events, including the release of neurotransmitters and other secretions, metabolic adjustments, cell proliferation, contraction, and control of gene expression. Voltage-gated calcium channels are heterooligomeric complexes composed of several structurally dissimilar subunits (1, 2, , , and ), with the large 1 subunit forming the ion-conducting pore of the channel (Catterall et al., 2005b). Several classes of calcium channels have been identified based on their biophysical, electrophysiological, and pharmacological properties. An important biophysical distinction is that certain channels are activated by only small depolarizations (i.e., low-voltage activated), whereas others are activated only in the presence of larger depolarizations (high-voltage activated). The high-voltage activated channels are further subdivided on the basis of tissue distribution and biophysical and pharmacological properties into five classes (L-, N-, Q-, P-, and R-types), whereas the low-voltage activated channels are represented by only one biophysical and pharmacological class (T-type). The different functional classes of calcium

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channels correspond to subsets of the 10 known calcium channel 1 subunit genes (Catterall et al., 2005b). Direct effects of pyrethroids on voltage-gated calcium channels have been explored as a possible explanation for the enhanced neurotransmitter release in the CNS that accompanies the CS syndrome of intoxication (Soderlund et al., 2002) (see also Section 77.3.4). Two experimental approaches have been employed: biochemical studies of neurotransmitter release and calcium uptake using presynaptic nerve terminals isolated from brain, and electrophysiological studies of native calcium currents in various cell types or of calcium currents carried by calcium channels of defined subunit structure expressed in Xenopus oocytes or HEK (human embryonic kidney) cells. Shafer and Meyer (2004) provide a comprehensive and critical review of data on the action of pyrethroids on voltagegated calcium channels published up to 2003. The following sections summarize this earlier work and review new information published since 2003.

77.8.1.1  Biochemical Studies The first in vitro findings that established the direct action of pyrethroids on the spontaneous release of neurotransmitters were obtained using isolated presynaptic nerve terminals (synaptosomes) from guinea pig cortex (Nicholson et al., 1987). Both deltamethrin and permethrin increased spontaneous release, but permethrin was much less potent than deltamethrin. The spontaneous release of neurotransmitter was only partially blocked by TTX and was partially dependent upon external calcium. A series of 25 pyrethroids also increased the spontaneous release of neurotransmitters from rat brain synaptosomes (Doherty et al., 1986a, 1987). Release promoted by pyrethroids was only partially abolished by TTX or by substituting choline for sodium, indicating an action at a site other than the sodium channel. Fenvalerate, cypermethrin, and deltamethrin also increased the spontaneous, calciumdependent release of dopamine and acetylcholine from rabbit striatal brain slices (Eells and Dubocovich, 1988). This spontaneous release was concentration dependent and was specific for the neurotoxic isomers of these pyrethroids. In the case of fenvalerate, spontaneous release was completely blocked by TTX, which tends to rule out a direct action on calcium channels as a mechanism for enhanced release at least in these striatal slice assays. Enhancement of pyrethroid-induced neurotransmitter rele­ ase is also evident following nerve terminal depolarization. Treatment with deltamethrin, cypermethrin, and fenvalerate greatly enhanced a calcium-dependent neurotransmitter rele­ase following the depolarization of rat brain synaptosomes by potassium (Brooks and Clark, 1987). The deltamethrininduced release was highly correlated with calcium uptake and only partially blocked by TTX (Clark and Brooks, 1989), but release was completely blocked by D595, a potent phenethylamine calcium channel blocker (Brooks and Clark, 1987).

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Recent studies (Symington and Clark, 2005; Symington et al., 2007b, 2008) provide a more comprehensive picture of the action of pyrethroids on both calcium uptake and depolarization-evoked neurotransmitter release in rat brain synaptosome preparations. These studies employed 11 commercially available pyrethroids that included members of both the Type I and Type II structural classes and compounds previously characterized as producing either the T or CS syndromes of intoxication. Five of the six Type II compounds (cyfluthrin, cyhalothrin, cypermethrin, deltamethrin, and esfenvalerate) and permethrin, a Type I compound, were potent enhancers of both calcium uptake and neurotransmitter release. For deltamethrin, both of these effects were inhibited by -conotoxin MVIIC, a selective blocker of N-type calcium channels. The remaining five compounds either were much less potent in both assays (cismethrin and bifenthrin) or enhanced glutamate release without a corresponding enhancement of calcium influx (bioallethrin, fenpropathrin, and tefluthrin). Since these experiments were performed in the presence of TTX to eliminate effects on voltage-gated sodium channels, the results provide clear evidence that at least some pyrethroids directly affect both calcium uptake and neurotransmitter release in isolated presynaptic terminals from brain. Efforts to identify interactions of pyrethroids with calcium channels using radioligands that label specific calcium channel binding domains have met with limited success. Several pyrethroids have been reported to inhibit the binding of a tritiated L-type calcium channel blocker, [3H]nimodipine, to rat brain synaptosomes (Ramadan et al., 1988a). However, the affinity for pyrethroids in this assay was low so that the concentration required for inhibition exceeded that necessary to modify voltage-sensitive sodium channels. In another study, deltamethrin displaced [3H]verapamil (a phenethylamine calcium channel blocker of L- and T-type calcium channels) binding to rat brain synaptic membranes (Kadous et al., 1994). It is likely that the failure to detect effects of pyrethroids in binding assays that correlate with effects on neurotransmitter release is due to the use of radioligands that bind selectively to calcium channel subtypes that are not involved in presynaptic neurotransmitter release.

77.8.1.2  Electrophysiological Studies Patch clamp techniques have been employed to detect distinct L- and T-type calcium currents in mouse neuroblastoma (N1E-115) cells. In these cells, the non-cyano pyrethroid tetramethrin blocked 75% of the T-type calcium current but only 30% of the L-type current (Yoshii et al., 1985). In contrast, the -cyano pyrethroids deltamethrin and fenvalerate had no effect on either calcium current. Similar results were obtained in assays of the T- and L-type calcium currents in single isolated sino-atrial node cells of rabbit heart. In this study, tetramethrin completely blocked only the T-type calcium current and resulted in a significant slowing of the diastolic

Chapter | 77  Toxicology and Mode of Action of Pyrethroid Insecticides

depolarization in the latter two-thirds of the pacemaker’s action potential in heart muscle (Hagiwara et al., 1988). In studies with intestinal smooth muscle cells from guinea pig, tetramethrin abolished the T-type calcium currents in a dosedependent manner but had no effect on L-type currents (Yabu et al., 1989). In neurosecretory cells, tefluthrin produced partial block of L-type calcium currents at concentrations that also caused extensive modification of the sodium currents in these cells (Wu et al., 2009). Recent studies have examined the effects of pyrethroids on single calcium channel isoforms expressed either in HEK cells or in Xenopus oocytes. Hildebrand et al. (2004) showed that allethrin was equally effective as a blocker of T-, L-, or P-/Q-type calcium channel isoforms expressed in HEK cells. For all three channel types, block was accompanied by an increase in the rate of channel inactivation and a shift of the voltage dependence of inactivation to more negative potentials. In addition, block of L- and P-/Q-type channels was enhanced by repeated stimulation, whereas block of T-type channels was not affected. Similarly, Symington and Clark (2005) showed that deltamethrin caused the partial block of an N-type calcium channel isoform expressed in Xenopus oocytes. In contrast to calcium channels expressed in HEK cells, pyrethroid block of N-type channels in oocytes was accompanied by a slight hyperpolarizing shift in the voltage dependence of activation and a slowing of channel activation and inactivation. Interestingly, insertion of the T422E mutation in the rat Cav2.2 (N-type) calcium channel reversed the effect of deltamethrin, which caused a significant increase in the peak transient calcium current in the mutated channel (Symington et al., 2007a). The T422E mutation mimics complete phosphorylation of threonine-422, suggesting that deltamethrin may differentially affect up- and downregulated calcium channels.

77.8.1.3  Toxicological Significance of Calcium Channel Effects Both biochemical and electrophysiological studies document direct effects of pyrethroids on calcium channel function in vitro, but a causal connection between these effects and pyrethroid intoxication remains unclear. Activation of nerve terminal calcium channels, an effect implied by the results of biochemical studies with synaptosomes, could conceivably contribute to the pronounced neurotransmitter release in the CNS that accompanies the CS intoxication syndrome. However, neurotransmitter release in vivo could also simply result from the prolonged activation of presynaptic sodium channels by Type II compounds that cause the CS syndrome. Channel block, the most consistently observed action of pyrethroids on calcium currents carried by different calcium channel subtypes, is an effect opposite that expected to cause or enhance neurotransmitter release. Further investigation is required to assess the differential sensitivity, if any, of different calcium channel subtypes to a wider structural variety of

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pyrethroids and to assess the broader significance of channel phosphorylation on pyrethroid action. More studies are also required to identify and define differential effects of pyrethroids on sodium and calcium currents in native neurons and other cell types that express both sodium and calcium channels.

77.8.2  Actions of Pyrethroids on   Voltage-Gated Chloride Channels Voltage-sensitive chloride channels are a structurally diverse and widely distributed group of cell membrane proteins (Gelband et al., 1996; Jentsch, 1996; Jentsch et al., 1999). Molecular cloning has revealed two structural classes of voltage-sensitive chloride channels that are distinct from each other and structurally unrelated to the voltage-sensitive cation (e.g., sodium and calcium) channel family (Jentsch, 1996). The CLC class of chloride channels, encoded by a multi-gene family, is involved in the stabilization of membrane resting potential, transepithelial transport, and cell volume regulation in virtually all cell types (Jentsch, 1996; Jentsch et al., 1999). A second structural class of voltage-dependent chloride channel proteins is exemplified by the cystic fibrosis transmembrane conductance regulator (CFTR), a cell membrane protein that regulates chloride conductance in epithelial cells that was first identified as the molecular locus of genetic defects that cause cystic fibrosis (Gelband et al., 1996; Jentsch, 1996). The pharmacology of voltage-dependent chloride channels is poorly characterized due to the lack of high-affinity ligands specific for these proteins (Gelband et al., 1996). Many compounds that affect or bind to voltage-dependent chloride channels, such as the convulsant t-butylbicyclophosphorotrithioate (TBPS), certain cyclodiene insecticides, avermectins, and barbiturates, also act at the GABA receptor–chloride ionophore complex (Abalis et al., 1985; Bloomquist, 1993b; Payne and Soderlund, 1991; Schwartz et al., 1984). The existence of pharmacological crosstalk between different chloride channel types complicates pharmacological discrimination between GABA-gated and voltage-sensitive chloride channels. Voltage-gated chloride channels have also come under consideration as alternative targets for pyrethroids that cause the CS intoxication syndrome. Investigation of the role of chloride channels has involved electrophysiological experiments in vitro coupled with pharmacological studies in vivo of the interactions between pyrethroids and agents known to act on chloride channels.

77.8.2.1  Electrophysiological Studies in Vitro Studies of the action of cismethrin and deltamethrin on skeletal muscle showed that deltamethrin but not cismethrin increased muscle membrane resistance, which was suggested to result from a block of the chloride permeability of

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the muscle membrane (Forshaw et al., 1987). A subsequent study (Forshaw and Ray, 1990) confirmed the effect of deltamethrin but not cismethrin on the input resistance of rat diaphragm skeletal muscle fibers and showed that a reduction in extracellular chloride ion concentration prevented the effects of deltamethrin. This study also documented similar effects of deltamethrin on the input resistance of rat vagus nerve preparations that were prevented by low extracellular chloride or treatment of the preparation with ivermectin, which activates neuronal voltage-dependent chloride channels (Abalis et al., 1986a). Patch clamp studies of single neuronal voltage-dependent chloride channels in excised membrane patches from N1E-115 neuroblastoma documented the blockade of single channel conductance by deltamethrin and cypermethrin but not cismethrin (Forshaw et al., 1993; Ray et al., 1997). This result suggested that chloride channel blockade might play a role in the action of pyrethroids that cause the CS intoxication syndrome. However, a structure–activity study (Burr and Ray, 2004) involving 14 pyrethroids encompassing compounds that produce both the T and CS syndromes failed to confirm this relationship. In this study esfenvalerate and -cyhalothrin, Type II structures expected to produce the CS intoxication syndrome, were without effect on chloride channels, whereas bioallethrin, a Type I structure that produces the T syndrome, blocked chloride channels.

77.8.2.2  Pharmacological Studies in Vivo Studies in vivo of the interactions between pyrethroids and agents known to act at voltage-sensitive chloride channels provide further insight into the involvement of this target in pyrethroid intoxication. These experiments involved coadministration of deltamethrin with ivermectin (which is known to activate voltage-sensitive chloride channels and have limited access to the CNS), pentobarbital (a barbiturate that selectively activates voltage-sensitive chloride channels), or phenobarbital (which exhibits sedative effects typical of barbiturates without activating voltage-sensitive chloride channels) (Forshaw et al., 2000; Ray et al., 1999). Intraperitoneal pretreatment of rats with ivermectin reduced the degree of salivation caused by subsequent intravenous treatment with deltamethrin and also reduced the incidence of deltamethrin mortality and the motor signs of deltamethrin intoxication at this dose, but ivermectin only affected salivation at a lower dose of deltamethrin (Forshaw et al., 2000; Ray et al., 1999). Ivermectin also reduced the severity of the direct effects of deltamethrin on skeletal muscle excitability in urethane-anesthetized rats (Forshaw et al., 2000; Ray et al., 1999). In parallel experiments, pentobarbital significantly antagonized the motor signs of intoxication and reduced the degree of mortality in deltamethrin-treated rats but was less effective than ivermectin in reducing salivation (Forshaw et al., 2000; Ray et al., 1999). In contrast, an equi-sedative

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dose of phenobarbital (which lacks the selectivity of pentobarbital for voltage-sensitive chloride channels) reduced the number of deaths in deltamethrin-treated rats but did not significantly affect salivation or the motor signs of intoxication (Forshaw et al., 2000; Ray et al., 1999). These studies were also extended to examine the effects of barbiturates on intoxication by cismethrin, a pyrethroid without demonstrable effects on voltage-sensitive chloride channels in vitro (Forshaw et al., 1987; Forshaw and Ray, 1990). Pentobarbital did not affect either the motor signs or the number of deaths in rats treated intravenously with cismethrin, but phenobarbital produced a significant reduction in lethality under these conditions (Ray et al., 1999). These authors concluded that the effects of ivermectin and pentobarbital on the signs of deltamethrin intoxication reflected a specific antagonism of the action of deltamethrin on voltage-sensitive chloride channels, whereas the effects of pentobarbital and phenobarbital on the lethality of both deltamethrin and cismethrin were ascribed to central effects on neuronal excitability that were not specifically attributable to an action on chloride channels (Ray et al., 1999).

77.8.2.3  Toxicological Significance of Chloride Channel Effects Evidence for the toxicological relevance of pyrethroid actions on voltage-sensitive chloride channels comes primarily from the results of pharmacological studies in vivo. Both ivermectin and pentobarbital, which selectively activate voltage-dependent chloride channels, antagonized the signs of deltamethrin intoxication and reduced the incidence of lethality at the deltamethrin doses employed but did not similarly affect cismethrin intoxication. The effects of these agents on salivation and the motor signs of intoxication caused by deltamethrin appeared to be due to antagonism of the action of deltamethrin at peripheral and CNS chloride channels, respectively. Despite the limited number of pyrethroids used in these studies, the results suggested that actions at voltage-sensitive chloride channels may contribute to some of the signs of intoxication associated with the CS poisoning syndrome. However, the extension of this interpretation to all compounds that produce the CS syndrome is limited by more recent studies showing that not all CS syndrome compounds block chloride channels in vitro. Thus, effects on chloride channels may contribute to the production of signs of intoxication associated with the CS syndrome for certain pyrethroids but are not likely to be the primary cause of these effects for all compounds (Burr and Ray, 2004).

77.8.3  Actions Of Pyrethroids on GABAA Receptors The GABA (-aminobutyric acid) receptor–chloride ionophore complex (GABAA receptor) is an important mediator

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of inhibitory neurotransmission in the mammalian nervous system (Macdonald and Olsen, 1994). Release of GABA by presynaptic nerve terminals activates a chloride channel on the postsynaptic membrane, leading to hyperpolarization of the postsynaptic nerve terminal and thus requiring an increase in the amount of excitatory input into that terminal in order to excite the postsynaptic neuron. GABA receptors are important target sites for the action of several classes of therapeutic drugs and toxicants (Johnston, 1996). Ethanol, barbiturates, and benzodiazepine anxiolytics allosterically modify the action of GABA. Blockers of the chloride channel, which include naturally occurring compounds such as picrotoxinin and synthetic insecticides such as the chlorinated cyclodienes (e.g., dieldrin, endosulfan) (Bloomquist, 1998) and phenylpyrazoles (e.g., fipronil) (Hosie et al., 1995), act as convulsants by blocking the inhibitory effects of GABA and producing an indirect augmentation of excitatory neurotransmission.

77.8.3.1  Biochemical Studies in Vitro The first report of an interaction of pyrethroids on GABA receptors showed the stereospecific inhibition by deltamethrin but not its nontoxic -R epimer of the binding of [3H]dihydropictoxinin to the convulsant (chloride channel) site of rat brain GABA receptors (Leeb-Lundberg and Olsen, 1980). The subsequent development of [35S]TBPS as an improved radioligand for the chloride channel site of GABA receptors led to the further documentation of the interaction of pyrethroids with the convulsant/chloride channel site (Lawrence and Casida, 1983; Lawrence et al., 1985). These studies, employing 37 pyrethroids, documented the inhibition of [35S]TBPS binding by the toxic isomers of four pyrethroids containing the -cyano-3-phenoxybenzyl moiety (cypermethrin, deltamethrin, fenvalerate, and fluvalinate) and by isomer mixtures of two other -cyano compounds (cyphenothrin and fenpropathrin) but not by the nontoxic isomers of -cyano compounds or by any pyrethroids lacking the cyano substituent. However, the original claims for absolute stereospecificity in these assays were contradicted by the subsequent finding that the 1S,cis,S isomer of cypermethrin, which has low acute toxicity to mammals, produced significant inhibition of [35S]TBPS binding to rat brain GABA receptors (Seifert and Casida, 1985). The structure–activity correlations for pyrethroid-dependent inhibition of [35S]TBPS binding led to the widely recognized hypothesis that -cyano pyrethroids caused the CS intoxication syndrome by an action at the GABA receptor–ionophore complex (Lawrence and Casida, 1983). Further analysis of the action of pyrethroids on GABA receptors was undertaken using assays of GABA receptor function. Several studies of the effects of pyrethroids on GABA-stimulated chloride-36 uptake into brain vesicles confirmed the action of neurotoxic isomers of -cyano pyrethroids as antagonists at mammalian brain GABA

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receptors (Abalis et al., 1986b; Bloomquist et al., 1986; Bloomquist and Soderlund, 1985; Ramadan et al., 1988b). However, the inhibition of chloride uptake in these studies was typically incomplete at maximally effective pyrethroid concentrations. Also, the incomplete stereoselectivity of pyrethroid action on GABA receptors in these assays was inconsistent with the profound stereospecificity of pyrethroid toxicity. For example, the enantiomer of deltamethrin, which is at least 500-fold less toxic than deltamethrin, was only 10-fold less effective as an inhibitor of GABA-dependent chloride uptake (Bloomquist et al., 1986; Bloomquist and Soderlund, 1985).

77.8.3.2  Electrophysiological Studies Electrophysiological assays have been employed to assess the relative sensitivity of GABA receptors and voltage-sensitive sodium channels expressed in the same cell or neuronal pathway to pyrethroids. GABA receptors in cultured dorsal root ganglion neurons were much less sensitive to the actions of deltamethrin than were the populations of voltage-sensitive sodium channels expressed in the same cells (Ogata et al., 1988). Also, electrophysiological recordings from defined GABAergic pathways in the rat hippocampus (Gilbert et al., 1989; Joy and Albertson, 1991; Joy et al., 1989, 1990) showed that the effects of pyrethroids were consistent with an augmentation of inhibition, possibly as a result of sodium channel-mediated presynaptic excitation of GABAergic neurons, rather than the antagonism of inhibition that is typical of established blockers of the GABA-gated chloride channel.

77.8.3.3  Toxicological Significance of GABAA Receptor Effects A large body of biochemical evidence documents the ability of Type II pyrethroids to bind to and block GABA receptors in mammalian brain preparations. Blockade of GABA receptors is an indirect neuroexcitatory effect, involving the removal of inhibitory neuronal input, and is the established mode of action for convulsants such as picrotoxinin. From a functional point of view, an action on GABA receptors is therefore consistent with the neuroexcitatory signs of pyrethroid intoxication in vivo. The action of pyrethroids on GABA receptors is somewhat stereoselective for neurotoxic isomers of -cyano compounds but does not exhibit the absolute stereospecificity predicted by structure–toxicity relationships. In experimental systems where effects on GABA receptors and sodium channels can be assayed in the same preparation, GABA receptor blockade is not observed at concentrations of pyrethroid that disrupt sodium channel function. There is only a limited amount of information correlating the in vitro biochemical effects of pyrethroids on GABA receptors with intoxication in vivo. The

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initial description of the T and CS intoxication syndromes following intracerebral administration of pyrethroids to mice noted that the CS syndrome was similar to the signs of intoxication produced by the intracerebral administration of picrotoxinin (Lawrence and Casida, 1982). The ability of diazepam, a benzodiazepine known to act at GABA receptors, to increase the latency of the CS syndrome and to antagonize the acute intracerebral toxicity of deltamethrin and the neurotoxic isomer of fenvalerate more than 10-fold was also interpreted as evidence for an action at GABA receptors in the production of the CS syndrome (Gammon et al., 1982). However, the relatively low potency and incomplete stereospecificity of pyrethroids as GABA receptor antagonists in functional assays do not support an action at the GABA receptor as a significant target site involved in the production of the CS intoxication syndrome.

Conclusion In the three decades since the introduction of the first compounds with sufficient photostability for agricultural use, synthetic pyrethroids have become important pest management tools in agriculture, public health, and a variety of household applications. Pyrethroids are neurotoxic not only to insects but also to mammals. For most pyrethroids, acute toxicity is low to moderate and is limited by inefficient absorption by some routes and by rapid metabolic detoxication. However, the intrinsic toxicity of pyrethoids, which is observed upon direct administration to the CNS, is quite high. Structural subclasses of pyrethroids produce two distinct syndromes of acute intoxication, which have complicated efforts to forge causal connections between effects on neuronal target sites in vitro and the physiological consequences of poisoning in vivo. The insecticidal properties of pyrethroids derive from their ability to alter the function of voltage-gated sodium channels in insect neuronal membranes, thereby disrupting electrical signaling in the nervous system. Although the discovery and development process for pyrethroids involved the optimization of insecticidal activity using whole insects, optimization for high intrinsic activity on insect sodium channels was implicit throughout this process. The structure, function, and pharmacology of voltage-gated sodium channels are highly conserved between insects and mammals. It is therefore not surprising that pyrethroids also alter the function of mammalian sodium channels. For structural subclasses of pyrethroids, qualitative differences in sodium channel modification are generally correlated with the production of different intoxication syndromes, suggesting that actions on sodium channels are sufficient to account for the acute toxicity of this insecticide class. Whereas there is broad agreement that voltage-gated sodium channels are the primary target sites for the neurotoxic actions of pyrethroids, they may not be the only targets

involved in pyrethroid intoxication in mammals. In addition to their effects on sodium channels, pyrethroids also are known to affect a variety of voltage- and ligand-gated ion channels. Some of these putative alternative targets have subsequently been shown to have little relevance to intoxication, but actions on voltage-gated calcium and chloride channels remain of interest and may contribute to the neurotoxic effects of at least some compounds.

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Ginsburg, K. S., and Narahashi, T. (1993). Differential sensitivity of tetrodotoxin-sensitive and tetrodotoxin-resistant sodium channels to the insecticide allethrin in rat dorsal root ganglion neurons. Brain Res. 627, 239–248. Glowa, J. R. (1986). Acute and sub-acute effects of deltamethrin and chlordimeform on schedule-controlled responding in the mouse. Neurobehav. Toxicol. Teratol. 8, 97–102. Goldin, A. L. (2001). Resurgence of sodium channel research. Annu. Rev. Physiol. 63, 871–894. Gotoh, Y. et al. (1998). Permethrin emulsion ingestion: clinical manifestations and clearance of isomers. Clin. Toxicol. 36, 57–61. Gray, A. J., and Soderlund, D. M. (1985). Mammalian toxicology of pyrethroids. In “Insecticides” (D. H. Hutson and T. R. Roberts, eds.) Vol. 5, pp. 193–248. Wiley, New York. Hagiwara, N. et al. (1988). Contribution of two types of calcium currents to the pacemaker potentials of rabbit sino-atrial node cells. J. Physiol. 395, 233–253. He, F. et al. (1989). Clinical manifestations and diagnosis of acute pyrethroid poisoning. Arch. Toxicol. 63, 54–58. Hijzen, T. H., and Slangen, J. L. (1988). Effects of type I and type II pyrethroids on the acoustic startle response in rats. Toxicol. Lett. 40, 141–152. Hildebrand, M. E. et al. (2004). Mammalian voltage-gated calcium channels are potently blocked by the pyrethroid insecticide allethrin. J. Pharmacol. Exp. Ther. 308, 805–813. Hosie, A. M. et al. (1995). Actions of the insecticide fipronil, on dieldrinsensitive and -resistant GABA receptors of Drosophila melanogaster. Br. J. of Pharmacol. 115, 909–912. Hudson, P. M. et al. (1986). Neurobehavioral effects of permethrin are associated with alterations in regional levels of biogenic amine metabolites and amino acid neurotransmitters. NeuroToxicology 7, 143–154. Husain, R. et al. (1996). Behavioral, neurochemical, and neuromorphological effects of deltamethrin in adult rats. J. Toxicol. Environ. Health 48, 515–526. Jentsch, T. J. (1996). Chloride channels: a molecular perspective. Curr. Opin. Neurobiol. 6, 303–310. Jentsch, T. J. et al. (1999). The CLC chloride channel family. Pflugers Arch. 437, 783–795. Johnston, G. A. R. (1996). GABAA receptor pharmacology. Pharmacol. Ther. 69, 173–198. Joy, R. M., and Albertson, T. E. (1991). Interactions of GABAA antagon­ists with deltamethrin, diazepam, pentobarbital, and SKF100330A in the rat dentate gyrus. Toxicol. Appl. Pharmacol. 109, 251–262. Joy, R. M. et al. (1989). Type I and Type II pyrethroids increase inhibition in the hippocampal dentate gyrus of the rat. Toxicol. Appl. Pharmacol. 98, 398–412. Joy, R. M. et al. (1990). Characteristics of the prolonged inhibition produced by a range of pyrethroids in the rat hippocampus. Toxicol. Appl. Pharmacol. 103, 528–538. Kadous, A. et al. (1994). High affinity binding of 3H-verapamil to rat brain synaptic membrane is antagonized by pyrethroid insecticides. J. Environ. Sci. Health 29B, 855–871. Lawrence, L. J., and Casida, J. E. (1982). Pyrethroid toxicology: mouse intracerebral structure-toxicity relationships. Pestic. Biochem. Physiol. 18, 9–14. Lawrence, L. J., and Casida, J. E. (1983). Stereospecific action of pyrethroid insecticides on the -aminobutyric acid receptor-ionophore complex. Science 221, 1399–1401. Lawrence, L. J. et al. (1985). Interactions of pyrethroid insecticides with chloride ionophore-associated binding sites. Neurotoxicology 6, 87–98.

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Lee, S. H., and Soderlund, D. M. (2001). The V410 M mutation associated with pyrethroid resistance in Heliothis virescens reduces the pyrethroid sensitivity of house fly sodium channels expressed in Xenopus oocytes. Insect Biochem. Mol. Biol. 31, 19–29. Leeb-Lundberg, F., and Olsen, R. W. (1980). Picrotoxinin binding as a probe of the GABA postsynaptic membrane receptor-ionophore complex. In “Psychopharmacology and Biochemistry of Neurotransmitter Receptors” (H. I. Yamamura et al., eds.), pp. 593–606. Elsevier, New York. Lessenger, J. E. (1992). Five office workers inadvertently exposed to cypermethrin. J. Toxicol. Environ. Health 35, 261–267. Lipkind, G. M., and Fozzard, H. A. (2000). KcsA crystal structure as framework for a molecular model of the Na channel pore. Biochemistry 39, 8161–8170. Litovitz, T. L. et al. (2002). 2001 annual report of the American Association of Poison Control Centers Toxic Exposure Surveillance System. Am. J. Emerg. Med. 20, 391–452. Lock, E. A., and Berry, P. N. (1981). Biochemical changes in the rat cerebellum following cypermethrin administration. Toxicol. Appl. Pharmacol. 59, 508–514. Lombet, A. et al. (1988). Interactions of insecticides of the pyrethroid family with specific binding sites on the voltage-dependent sodium channel from mammalian brain. Brain Res. 459, 44–53. Lund, A. E., and Narahashi, T. (1983). Kinetics of sodium channel modification as the basis for the variation in the nerve membrane effects of pyrethroids and DDT analogs. Pestic. Biochem. Physiol. 20, 203–216. Macdonald, R. L., and Olsen, R. W. (1994). GABAA receptor channels. Annu. Rev. Neurosci. 17, 569–602. Meacham, C. A. et al. (2008). Developmentally-regulated sodium channel subunits are differentially sensitive to -cyano containing pyrethroids. Toxicol. Appl. Pharmacol., 273–281. Meadows, L. S., and Isom, L. L. (2005). Sodium channels as macromolecular complexes: implications for inherited arrhythmia syndromes. Cardiovasc. Res. 67, 448–458. Moser, V. C. (1989). Screening approaches to neurotoxicity: a functional observational battery. J. Am. Coll. Toxicol. 8, 85–93. Muhammad, B. Y. et al. (2003). Developmental toxicity of pyrethroids. Arch. Toxicol. 77, 48–49. Narahashi, T. (1992). Nerve membrane Na channels as targets of insecticides. Trends Pharmacol. Sci. 13, 236–241. Narahashi, T. (1996). Neuronal ion channels as the target sites of insecticides. Pharmacol. Toxicol. 78, 1–14. Nicholson, R. A. et al. (1987). Pyrethroid- and DDT-evoked release of GABA from the nervous system in vitro. In “Pesticide Chemistry: Human Welfare and the Environment” (J. Miyamoto and P. C. Kearney, eds.) Vol. 3, pp. 75–78. Pergamon, Oxford. O’Reilly, A. O. et al. (2006). Modelling insecticide binding sites at the voltage-gated sodium channel. Biochem. J. 396, 255–263. Ogata, N. et al. (1988). Lindane but not deltamethrin blocks a component of GABA-activated chloride channels. Fed. Am. Soc. Exp. Biol. J. 2, 2895–2900. Payne, G. T., and Soderlund, D. M. (1991). Activation of -aminobutyric acid insensitive chloride channels in mouse brain synaptic vesicles by avermectin B1a. J. Biochem. Toxicol. 6, 283–292. Peele, D. B., and Crofton, K. M. (1987). Pyrethroid effects on schedulecontrolled behavior: time and dosage relationships. Neurotoxicol. Teratol. 9, 387–394. Pickett, J. A. (2004). New opportunities in neuroscience, but a great danger that some may be lost. In “Neurotox ‘03: Neurotoxicological Targets from Functional Genomics and Proteomics” (D. J. Beadle et al�������������������������������������������������������������������������������������������������������� .������������������������������������������������������������������������������������������������������� , eds.), pp. 1–10. Society of Chemical Industry, London.

Chapter | 77  Toxicology and Mode of Action of Pyrethroid Insecticides

Ramadan, A. A. et al. (1988a). Actions of pyrethroids on K-stimulated calcium uptake by, and [H]nimodipine binding to, rat brain synaptosomes. Pestic. Biochem. Physiol. 32, 114–122. Ramadan, A. A. et al. (1988b). Action of pyrethroids on GABAA receptor function. Pestic. Biochem. Physiol. 32, 97–105. Ray, D. E. et al. (1999). A new basis for therapy against Type-II pyrethroid poisoning. In “Progress in Neuropharmacology and Neurotoxicology of Pesticides and Drugs” (D. J. Beadle ed.), pp. 204–214. Royal Society of Chemistry, Cambridge. Ray, D. E. et al. (1997). Action of pyrethroid insecticides on voltagegated chloride channels in neuroblastoma cells. NeuroToxicology 18, 755–760. Rizzo, M. A. et al. (1996). Mechanisms of paresthesiae, dysthesiae, and hyperesthesiae: role of Na channel heterogeneity. Eur. Neurol. 36, 3–12. Ruzo, L. O. et al. (1979). Decamethrin metabolites from oxidative, hydrolytic and conjugative reactions in mice. J. Agric. Food Chem. 27, 725–731. Salgado, V. L., and Narahashi, T. (1993). Immobilization of sodium channel gating charge in crayfish giant axons by the insecticide fenvalerate. Mol. Pharmacol. 43, 626–634. Sattelle, D. B., and Yamamoto, D. (1988). Molecular targets of pyrethroid insecticides. Adv. Insect Physiol. 20, 147–213. Schwartz, R. D. et al. (1984). Barbiturate and picrotoxin-sensitive chloride efflux in rat cerebral cortical synaptoneurosomes. FEBS Lett. 175, 193–196. Seifert, J., and Casida, J. E. (1985). Solubilization and detergent effects on interactions of some drugs and insecticides with the t-butylbicyclophosphorothionate binding site within the -aminobutyric acid receptor-ionophore complex. J. Neurochem. 44, 110–116. Shafer, T. J., and Meyer, D. A. (2004). Effects of pyrethroids on voltagesensitive calcium channels: a critical evaluation of strengths, weaknesses, data needs, and relationship to assessment of cumulative neurotoxicity. Toxicol. Appl. Pharmacol. 196, 303–318. Shafer, T. J. et al. (2005). Developmental neurotoxicity of pyrethroid insecticides: critical review and future research needs. Environ. Health Perspect. 113, 123–136. Shafer, T. J. et al. (2008). Complete inhibition of spontaneous activity in neuronal networks in vitro by deltamethrin and permethrin. Neurotoxicology 29, 203–212. Sheets, L. P. (2000). A consideration of age dependent differences in susceptibility to organophosphorus and pyrethroid insecticides. NeuroToxicology 21, 57–64. Sheets, L. P. et al. (1994). Age-dependent differences in the susceptibility of rats to deltamethrin. Toxicol. Appl. Pharmacol. 126, 186–190. Shrivastava, I. H. et al. (2004). A model of voltage gating developed using the KvAP channel crystal structure. Biophys. J. 87, 2255–2270. Smith, T. J. et al. (1998). Actions of the pyrethroid insecticides cismethrin and cypermethrin on house fly Vssc1 sodium channels expressed in Xenopus oocytes. Arch. Insect Biochem. Physiol. 38, 126–136. Smith, T. J., and Soderlund, D. M. (1998). Action of the pyrethroid insecticide cypermethrin on rat brain IIa sodium channels expressed in Xenopus oocytes. NeuroToxicology 19, 823–832. Smith, T. J., and Soderlund, D. M. (2001). Potent actions of the pyrethroid insecticides cismethrin and cypermethrin on rat tetrodotoxin-resistant peripheral nerve (SNS/PN3) sodium channels expressed in Xenopus oocytes. Pestic. Biochem. Physiol. 70, 52–61. Soderlund, D. M. (1995). Mode of action of pyrethrins and pyrethroids. In “Pyrethrum Flowers: Production, Chemistry, Toxicology, and Uses” (J. E. Casida and G. B. Quistad, eds.), pp. 217–233. Oxford University Press, New York.

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Soderlund, D. M. (2005). Sodium channels. In “Comprehensive Molecular Insect Science” (L. Gilbert et al., eds.) Vol. 5, pp. 1–24. Elsevier, New York. Soderlund, D. M., and Bloomquist, J. R. (1989). Neurotoxic actions of pyrethroid insecticides. Annu. Rev. Entomol. 34, 77–96. Soderlund, D. M. et al. (2002). Mechanisms of pyrethroid toxicity: implications for cumulative risk assessment. Toxicology 171, 3–59. Soderlund, D. M., and Knipple, D. C. (2003). The molecular biology of knockdown resistance to pyrethroid insecticides. Insect Biochem. Mol. Biol. 33, 563–577. Soderlund, D. M., and Lee, S. H. (2001). Point mutations in homology domain II modify the sensitivity of rat Nav1.8 sodium channels to the pyrethroid cismethrin. Neurotoxicology 22, 755–765. Song, J.-H. et al. (1996). Interactions of tetramethrin, fenvalerate and DDT at the sodium channel in rat dorsal root ganglion neurons. Brain Res. 708, 29–37. Song, J.-H., and Narahashi, T. (1996). Modulation of sodium channels of rat cerebellar Purkinje neurons by the pyrethroid tetramethrin. J. Pharmacol. Exp. Ther. 277, 445–453. Stein, E. A. et al. (1987). Effects of pyrethroid insecticides on operant responding maintained by food. Neurotoxicol. Teratol. 9, 27–31. Symington, S. B., and Clark, J. M. (2005). Action of deltamethrin on Ntype (Cav2.2) voltage-sensitive calcium channels in rat brain. Pestic. Biochem. Physiol. 82, 1–15. Symington, S. B. et al. (2008). Characterization of 11 commercial pyrethroids on the functional attributes of rat brain synaptosomes. Pestic. Biochem. Physiol. 92, 61–69. Symington, S. B. et al. (2007a). Mutation of threonine-422 to glutamic acid mmics the phosphorylation state and alters the action of deltamethrin on Cav2.2. Pestic. Biochem. Physiol. 88, 312–320. Symington, S. B. et al. (2007b). Action of cismethrin and deltamethrin on functional attributes of isolated presynaptic nerve terminals from rat brain. Pestic. Biochem. Physiol. 87, 172–181. Tabarean, I. V., and Narahashi, T. (1998). Potent modulation of tetrodotoxin-sensitive and tetrodotoxin-resistant sodium channels by the Type II pyrethroid deltamethrin. J. Pharmacol. Exp. Ther. 284, 958–965. Tan, J. et al. (2005). Identification of amino acid residues in the insect sodium channel critical for pyrethroid binding. Mol. Pharmacol. 67, 513–522. Tan, J. et al. (2008). Action of pyrethroid insecticides on rat Nav1.6 sodium channels expressed in Xenopus oocytes. Toxicologist 97 Abstract 2274.. Tan, J., and Soderlund, D. M. (2009a). Actions of S-bioallethrin, tefluthrin and deltamethrin on rat Nav1.6 sodium channels expressed in Xenopus oocytes. Toxicol. Appl. Pharmacol. (in preparation). Tan, J., and Soderlund, D. M. (2009b). Human and rat Nav1.3 voltagegated sodium channels differ in inactivation properties and sensitivity to the pyrethroid insecticide tefluthrin. Neurotoxicology 30, 81–89. Tatebayashi, H., and Narahashi, T. (1994). Differential mechanism of action of the pyrethroid tetramethrin on tetrodotoxin-sensitive and tetrodotoxin-resistant sodium channels. J. Pharmacol. Exp. Ther. 270, 595–603. Trainer, V. L. et al. (1997). High affinity binding of pyrethroids to the  subunit of brain sodium channels. Mol. Pharmacol. 51, 651–657. Trainer, V. L. et al. (1993). Neurotoxin binding and allosteric modulation at receptor sites 2 and 5 on purified and reconstituted rat brain sodium channels. J. Biol. Chem. 268, 17114–17119. Vais, H. et al. (2000a). A single amino acid change makes a rat neuronal sodium channel highly sensitive to pyrethroid insecticides. FEBS Lett. 470, 135–138.

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Vais, H. et al. (2000b). Activation of Drosophila sodium channels promotes modification by deltamethrin: reductions in affinity caused by knock-down resistance mutations. J. Gen. Physiol. 115, 305–318. Vais, H. et al. (2003). Mutations of the para sodium channel of Drosophila melanogaster identify putative binding sites for pyrethroids. Mol. Pharmacol. 64, 914–922. Verschoyle, R. D., and Aldridge, W. N. (1980). Structure-activity relationships of some pyrethroids in rats. Arch. Toxicol. 45, 325–329. Verschoyle, R. D., and Barnes, J. M. (1972). Toxicity of natural and synthetic pyrethrins to rats. Pestic. Biochem. Physiol. 2, 308–311. Vijverberg, H. P. M., and van den Bercken, J. (1990). Neurotoxicological effects and the mode of action of pyrethroid insecticides. CRC Crit. Rev. Toxicol. 21, 105–126. Watson, T. F. et al. (2004). 2003 annual report of the American Association of Poison Control Centers Toxic Exposure Surveillance System. Am. J. Emerg. Med. 22, 335–404. Watson, W. A. et al. (2005). 2004 annual report of the American Association of Poison Control Centers Toxic Exposure Surveillance System. Am. J. Emerg. Med. 23, 589–666. Watson, W. A. et al. (2003). 2002 annual report of the American Association of Poison Control Centers Toxic Exposure Surveillance System. Am. J. Emerg. Med. 21, 353–421.

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Whitaker, W. R. J. et al. (2001). Comparative distribution of voltage-gated sodium channel proteins in human brain. Mol. Brain Res. 88, 37–53. Wolansky, M. J. et al. (2006). Relative potencies for acute effects of pyrethroids on motor function in rats. Toxicol. Sci. 89, 271–277. Wolansky, M. J., and Harrill, J. A. (2008). Neurobehavioral toxicology of pyrethroid insecticides in adult animals: a critical review. Neurotoxicol. Teratol. 30, 55–78. Wu, S.-N. et al. (2009). Underlying mechanism of action of tefluthrin, a pyrethroid insecticide, on voltage-gated ion currents and on action currents in pituitary tumor (GH3) cells and GnRH-secreting (GT1-7) neurons. Toxicology 2009, 70–77. Yabu, H. et al. (1989). Two types of Ca channels in smooth muscle cells isolated from guinea- pig taenia coli. Adv. Exp. Med. Biol. 255, 129–134. Yoshii, M. et al. (1985). Effects of pyrethroids and veratridine on two types of calcium channels in neuroblastoma cells. Soc. Neurosci. Abstr. 11, 158.9. Zhao, Y. et al. (2004). A gating hinge in Na channels: a molecular switch for electrical signaling. Neuron 41, 859–865.

Section XII

Herbicides

(c) 2011 Elsevier Inc. All Rights Reserved.

Chapter 78

Dialkyldithiocarbamates (EBDCs) Susan Hurt1, Janet Ollinger1, Gail Arce2, Quang Bui3, Abraham J. Tobia4 and Bennard van Ravenswaay5 1

Rohm and Haas Company; 2Griffin LLC; 3Cerexagri, Inc.; 4Aventis; 5BASF AG

78.1  Chemistry and formulations Ethylenebisdithiocarbamates (EBDCs) are a group of fungicides that have been used widely throughout the world since the 1940s to protect a wide variety of crops against fungal disease. There are five members of the class, specifically mancozeb, maneb, metiram, zineb, and nabam. All members have an ethylenebisdithiocarbamate backbone, with different metals associated with the individual compounds. The structure of each compound is shown below. Mancozeb: [-MnSC(:S)NHCH2CH2NHC(:S)S-]xZny, where x/y  11 Maneb: [-MnSC(:S)NHCH2CH2NHC(:S)S-]x Metiram: [ [-(NH 3 )Zn-S-C(:S)NHCH 2 CH 2 NHC(:S)S-] 3 -S-C(:S)NHCH2CH2NHC(:S)S-]x Zineb: [-ZnSC(:S)NHCH2CH2NHC(:S)S-]x Nabam: [NaSC(:S)NHCH2CH2NHC(:S)SNa] The molecular weights of the individual EBDCs are mancozeb—271, maneb—265, metiram—1088.6, zineb—275, and nabam—256. At this time mancozeb, maneb, and metiram are the most widely used EBDCs. Zineb is used to a lesser degree and nabam is no longer used in agriculture. Thus, this chapter will focus on mancozeb, maneb, and metiram. The EBDCs are sold as wettable powder, dry flowable (also called water dispersable granules), and flowable formulations. EBDCs can also be sold as a premix with various blending partners.

78.2  Uses EBDCs are used to control about 400 fungal pathogens on more than 100 crops. The major EBDC uses around the world include grapes (fresh grapes, grapes grown for Hayes’ Handbook of Pesticide Toxicology Copyright © 2001 Elsevier Inc. All rights reserved

juice, and grapes grown for wine), potatoes, citrus, apples, tomatoes, melons, and bananas. EBDCs are also important products for disease control in corn, cereal grains, leafy vegetables, brassica vegetables, cranberries, onions, peanuts, sugar beets, asparagus, and nuts as well as for many other critical crops that are grown on a lower amount of acreage. Diseases of turf and ornamental crops are also controlled by EBDCs. Some of the economically important diseases controlled include early and late blight, downy mildew, and bacterial diseases. EBDCs are key components of fungicide resistance management programs because they have a multisite mode of action. For example, EBDCs deactivate the sulfhydryl containing enzymes which mediate numerous biosynthesis, mechanical, and transport activities within the fungal cytoplasm. They also inactivate ATP production, the Krebs cycle, the enzymes which convert glucose to pyruvate, and enzymes which convert amino and fatty acids to acetylcoenzyme A. Thus, resistance will not develop. After over 40 years of use no resistance has developed to any of the EBDCs.

78.3  Hazard identification The toxicology database supporting the assessment of the potential health risks of the EBDCs and their common metabolite ethylenethiourea (ETU) has been upgraded in recent years with a complete set of modern studies of mancozeb, maneb, and metiram conducted in full compliance with OECD and other applicable national and international guidelines and internationally recognized good laboratory practices. These newer studies have superseded the older studies in the published literature and now form the core of the toxicology database relevant for the hazard identification and dose-response assessment of this family of fungicides. 1689

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78.3.1  Pharmacokinetics and Metabolism Studies of the pharmacokinetics and metabolism of mancozeb, maneb, and metiram in laboratory animals have indicated that the EBDCs are only partially absorbed, then rapidly metabolized and excreted with no evidence of long-term bioaccumulation. Absorption of oral doses is rapid. Most of the administered dose is excreted within 24 hours, with about half eliminated in the urine and half in the feces. Biliary excretion is minimal, indicating that only approximately 50% of oral doses are absorbed. Only low level residues are found in tissues, principally in the thyroid. ETU is the major metabolite. On average 7.5% of an EBDC dose administered to rats is metabolized to ETU on a weight basis. The bioconversion factor in mice is slightly

smaller at 5 to 6% (Cameron et al., 1990; DiDonato and Longacre, 1986; Emmerling, 1978b; EPA, 1992; Hawkins et al., 1985; Kocialski, 1989; Longacre, 1986; Nelson, 1986, 1987; Piccirillo et al., 1992; Puhl, 1985). The spectrum of metabolites produced in laboratory animals points to two common metabolic pathways (Fig. 78.1), which both lead ultimately to the formation of glycine and incorporation into natural products. In the predominant pathway quantitatively, the dithiocarbamate linkages are hydrolyzed to produce ethylenediamine (EDA) directly, and EDA is oxidized to glycine, joining the intermediary metabolic pool at this point. The other pathway is responsible for the toxic effects of the EBDCs and involves oxidation to ethylenebisisothiocyanatesulfide and then to ETU, various derivatives of ETU, and ethyleneurea

S H N

S N

S− S− N H

S

S

S

N

H N

EBIS

EBDC

S N

N

Jaffe’s base

S HN

NH

ETU

CH3CONH

O

NH2 HN

NH

N-acetyl EDA ETU HOCHN

NHCOH

N, N’-formyl EDA HOCHN

HN

NH

Ethylene diamine

NH2

N-formyl EDA

Interconversion to other amino acids Proteins Nucleic acids Figure 78.1  Metabolic pathway of EBDC.

NH2CH2COOH Glycine

Pyruvate Acetyl Co-A Citrate

Sugars Saccharides Fatty acids Lipids

NH

Chapter | 78  Dialkyldithiocarbamates (EBDCs)

(EU) before rejoining the main pathway with conversion to EDA, glycine, and other natural products. ETU metabolism has also been extensively studied in multiple species. As with the EBDCs, oral doses are rapidly absorbed and rapidly excreted, although in this case primarily in the urine and more quickly in mice than in rats. In most species the greater majority (70% or more) of an oral dose is eliminated via the urine within 48 hours. Concentrations in blood and tissues are generally at comparably low levels with the exception of somewhat higher levels in the thyroid; levels in maternal and fetal tissues were similar 3 hours after dosing. Half-lives for elimination from maternal blood were 5.5 and 9.4 hours in mice and rats, respectively. Unchanged ETU was the principal metabolite in rats and guinea pigs, with small amounts of EU. In mice the principal identified metabolites were ETU and imidazolinylsulfenate, and in cats, S-methyl ETU was the principal metabolite (DiDonato and Longacre, 1987; Emmerling, 1978b; Iverson et al., 1977, 1980; Jordan and Neal, 1979; Kato et al., 1976; Peters et al., 1982; Ruddick et al., 1976a, 1977; Teshima et al., 1981). Studies of dermal absorption of the EBDCs have been challenging due to the difficulty of small scale preparation of radiolabelled samples representative of their ­complex polymeric structures. Thus, the reported values of 0.2–6.5% are considered to be overestimates of their actual dermal absorption potential under conditions of use (Craine, 1991; Haines, 1980; Hawkins et al., 1984; Tomlinson and Longacre, 1988). Dermal absorption of ETU increased from 5 to 22% with decreasing applied skin concentrations (DiDonato and Longacre, 1987).

78.3.2  Acute Toxicity The EBDC’s have very low acute toxicity by the oral, dermal, and respiratory routes (Table 78.1). The World Health Organization (WHO) has classified mancozeb, maneb, and metiram as unlikely to present an acute exposure hazard under conditions of normal use (WHO, 1994). Although not irritating to skin on initial contact and only slightly irritating to eyes and mucous membranes, prolonged or repeated skin contact may result in dermatitis due to their weak sensitization potential. ETU is only slightly toxic after oral administration, but it is a moderate to weak ­sensitizer in the guinea pig maximization test (Matsushita et al., 1976, 1977).

78.3.3  Short- and Long-Term Toxicity and Oncogenic Potential The EBDCs share a comparable toxicological profile, primarily based on the toxic effects of their common ETU metabolite. A summary of the critical doses and effects in subchronic and longer term studies is presented for each

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of the EBDCs in Table 78.2, and a similar summary for ETU is presented in Table 78.3. Results are collected by study type. As is illustrated repeatedly in these tables, the principal target organ upon repeated exposure to all of the EBDCs is the thyroid, which is also the principal target organ of ETU. For example, all three EBDCs (mancozeb, maneb, and metiram) and ETU altered thyroid hormone levels and/or weights at the lowest affected dose after three months of dietary feeding in rats. Most of the other organs affected, such as the liver at generally higher doses or red blood cells usually in dogs, are also common to ETU. Prolonged dietary feeding of ETU produces thyroid and pituitary tumors in rats and mice, and liver tumors in mice. As normally occurs with toxicological effects due to formation of a metabolite, the effects of ETU are not as strong in the EBDCs. When EBDCs are administered, much higher doses are required to produce adverse effects, and the effects themselves are generally not as pronounced or may be precluded altogether by high dose limitations.

78.3.3.1  Thyroid Effects The effects of the EBDCs on the thyroid after either shortor long-term dietary administration are consistent on both a quantitative and a qualitative basis with those of ETU. As described below, it is well accepted that these effects are the result of a secondary mechanism (hormonal imbalance), and that there is a threshold for the resulting tumors Similarly to the structurally related thionamide drugs (propylthiouracil, methimazole, and carbimazole) which are used clinically for treatment of hyperthyroidism in humans, the primary toxicological finding with ETU in laboratory animals is inhibition of the synthesis of thyroid hormones, thyroxine (T4) and triiodothyronine (T3), leading to elevated serum levels of thyroid stimulating hormone (TSH) via feedback stimulation of the hypothalamus and pituitary (Atterwill and Aylard, 1995; Engler and Burger, 1984; O’Neil and Marshall, 1984)(see Fig. 78.2). Prolonged and continuous elevation of serum TSH levels results in hypertrophy and hyperplasia of the thyroid follicular cells in rats, mice, hamsters, monkeys, and dogs (Briffaux, 1991, 1992; Gak et al., 1976; Chhabra et al., 1992; Freudenthal et al., 1977a; Graham and Hansen, 1972; Graham et al., 1973, 1975; Leber et al., 1978; O’Hara and DiDonato, 1985; Schmid et al., 1992; Ulland et al., 1972), and ultimately in the development of follicular nodular hyperplasia, adenoma, and/or carcinoma in rats and mice (Chhabra et al., 1992; Graham et al., 1973,1975; Schmid et al., 1992; Ulland et al., 1972), but not in hamsters (Gak et al., 1976). There is evidence for reversibility of the thyroid effects (Arnold et al., 1983). The mechanism of the crucial early steps of thyroid tumor formation by ETU is well understood. ETU reversibly inhibits thyroid peroxidase-catalyzed iodination and coupling of tyrosine residues into the thyroid hormone

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Table 78.1  Acute Toxicity of Mancozeb, Maneb, Metiram, and ETU Type of study acute

Active ingredient

Strain-sex

LD(LC)50 mg/kg bw Reference (mg/L)

Acute oral, rat

mancozeb

F344, M

5000

Watts and Chan (1984a,b)

mancozeb

CRCD, M

5000

DeCrescente and Parsons (1980)

maneb

Crl:CD BR, M/F

5000

Naas (1989a)

metiram

SD, M/F

6500–10,000

Jackh (1981); Leuschner (1979a); Hofmann (1985); Hofmann (1975)

Acute oral, mouse

mancozeb

B6C3Fi, M

5000

Watts and Chan (1984b)

Intraperitoneal, rat

mancozeb

Wistar, M/F

380

DeGroot (1974)

metiram

SD, M/F

318

Hofmann and Munk (1975)

Intraperitoneal, mouse

metiram

NMRI, M/F

80–215

Hofmann (1974); Leuschner (1979b)

Acute dermal, rat

metiram

SD, M/F

2000

Grundler (1979)

Acute dermal, rabbit

mancozeb

NZW, M

5000

DeCrescente and Parsons (1980)

maneb

NZW, M/F

2000

Naas (1989b)

mancozeb

COBS-CR (SD) BR, M/F

5.14 mg/fl

Hagan and Baldwin (1982)

maneb

Crl:CD BR, M/F

7.38 mg/l

Terrill (1990)

metiram

SD, M/F

5.7 mg/l

Klimisch and Zeller (1980)

ETU

M/F

545–ca. 2400

Peters et al. (1980a); Graham and Hansen (1972); Lewerenz and Plass (1984); Teramoto et al. (1978)

ETU

F (13 days pregnant)

600

Khera (1987)

ETU

M/F

ca. 2400–4000

Lewerenz and Plass (1984); Teramoto et al. (1978); Peters et al. (1980b)

ETU

F (9 days pregnant)

3000

Khera (1987)

ETU

F

3000

Teramoto et al. (1978)

ETU

F (11 days pregnant)

2400

Khera (1987)

Acute inhalation (4 hr)

Acute oral, rat

Acute oral, mouse

Acute oral, hamster

precursor thyroglobulin in vivo (Hill et al., 1989). Direct evidence for inhibition of thyroid hormone synthesis by ETU has been obtained in rats in vivo (Arnold et al., 1983; O’Neil and Marshall, 1984). ETU also reversibly inhibited thyroid peroxidase-catalyzed iodination reactions in vitro (Doerge and Takazawa, 1990). The correlation of the hormonal changes with hyperplasia and neoplasia has been clearly demonstrated in specific studies of ETU (Chhabra et al., 1992; Freudenthal et al., 1977a).

Similarly, the long-term stimulation of the pituitary via hypothalamic thyrotropin-releasing hormone also results in morphologic changes in the pituitary of rats, mice, monkeys, and dogs (Briffaux, 1991; Chhabra et al., 1992; Leber et al., 1978; Schmid et al., 1992), culminating in adenomas of the pars distalis after two years exposure in mice and rats (Chhabra et al., 1992; Schmid et al., 1992). A regenerative, nonhemolytic anemia that was observed in dogs is considered to be a secondary manifestation of

Chapter | 78  Dialkyldithiocarbamates (EBDCs)

1693

Table 78.2  EBDCs: Critical Findings in the Most Relevant Studies Study

Active

NOAEL (mg/kg/d)

LOAEL (mg/kg/d)

Effects observed at LOAEL/ critical results

Reference

Mouse 3 month oral

mancozeb

18

180

Decreased body weight and liver MFO activity; thyroid follicular hypertrophy and hyperplasia

O’Hara and DiDonato (1985)

metiram

84

302

Decreased T4

Gelbke et al. (1992a)

mancozeb

7.4

15

Decreased T4, increased TSH

Goldman et al. (1986)

maneb

5

25

Increased thyroid weight, follicular hyperplasia

Trutter (1988a)

metiram

6

20

Decreased T4 and iodine uptake, increased thyroid weight; microscopic changes in muscle fibers

Hunter et al. (1977)

5.8

23.2

Slight anemia; decreased T4, increased thyroid weight, general muscle weakness/ataxia and reduced grip strength without histopathological correlate at 70 mg/kg body weight/d

Gelbke et al. (1992b)

Rat 3 month oral

Rat 1 month dermal

mancozeb

1000

none

No systemic toxicity

Trutter (1988c)

Rabbit 1 month dermal

maneb

100

300

Increased follicular colloid

Trutter (1988b)

metiram

250

none

No systemic toxicity

Ullman et al. (1987)

mancozeb

79 mg/m3   6 hr/d ( 8.3)

326 mg/m3   6 hr ( 34)

Decreased weight gain, T4 levels; thyroid hyperplasia. All effects reversible after 13 weeks recovery

Hagan and Baldwin (1986); Hagan et al. (1986)

maneb

10 mg/m3  6 hr/d

30 mg/m3  6 hr/d

Decreased weight gain; all effects reversible after 13 weeks recovery

Ulrich (1986a, 1987)

metiram

2 mg/m3   6 hr/d

20 mg/m3   6 hr/d

Decreased weight gain; alveolar macrophage accumulation due to nonspecific dust reaction

Ulrich (1986b)

mancozeb

3

30

Decreased weight gain, RBC parameters

Cox (1986)

maneb

3.7

15

Thyroid follicular hyperplasia

Allen et al. (1989)

mancozeb

1.7

10.2

Decreased weight gain; no effect on reproduction below adult toxic levels

Muller (1992)

7

70

Decreased weight gain, feed consumption; increased thyroid, liver, kidney weights, microscopic changes in thyroid, kidney, and pituitary; no reproductive effect.

Solomon et al. (1988)

Rat 3 month inhalation

Dog 3 month

Rat two generation

(Continued )

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1694

Table 78.2  (Continued) Study

Active

NOAEL (mg/kg/d)

LOAEL (mg/kg/d)

Effects observed at LOAEL/ critical results

Reference

maneb

5.6

22.4

Increased liver, kidney weight ratios, thyroid follicular hyperplasia; no reproductive effect below toxic levels

Ryle et al. (1991)

metiram

1.8

14.4

Decreased body weight, feed consumption; no reproductive effect

Cozens et al. (1981)

Rat acute neurotoxicity

maneb

2000

NA

No adverse effect

Nemec (1993)

Rat 90-day neuropathology

mancozeb

8.2

48

Decreased feed consumption; neurohistopathological changes

Stadler (1991)

Rat developmental

mancozeb

32

128

Decreased maternal weight gain, feed consumption; teratogenic NOAEL; teratogenicity at 512 mg/kg body weight/d

Gallo et al. (1980)

60

360

Decreased maternal weight gain, feed consumption; maternal “reeling gait” and hindlimb paralysis, embryofetotoxicity

Tesh et al. (1988)

20

100

Decreased maternal weight gain, feed consumption; embryofetotoxicity

Nemec (1992)

100

500

Decreased maternal weight gain, feed consumption; hindlimb paresis; embryofetotoxicity and teratogenicity

Kapp et al. (1991)

metiram

80

160

Decreased maternal weight gain, slight decreases in litter size and weight

Palmer and Simons (1979)

mancozeb

55

100

Maternal weight loss, decreased feed consumption, increased abortions; no adverse embryofetal effects

Muller (1991)

30

80

Decreased maternal weight gain, feed consumption, litters; increased abortions, clinical signs, and deaths; no adverse embryofetal effects

Solomon and Holz (1987); Solomon and Lutz (1987)

metiram

10

40

Decreased maternal body weight, feed consumption: increased abortions; no adverse embryofetal effects

Gelbke et al. (1988)

mancozeb

7

28

Decreased weight gain, RBC parameters, inc, cholesterol

Shaw (1990)

2.3

23

Decreased weight gain, feed consumption, T4

Broadmeadow (1991a, b)

maneb

6.4

32

Thyroid thickening and enlargement, follicular hyperplasia

Corney et al. (1992)

metiram

2.5

31

Decreased T4, increased thyroid size and follicular hyperplasia, focal hepatic lipofuscin deposition, slight anemia, diarrhea, and blood biochemical changes

Corney et al. (1991)

maneb

Rabbit developmental

Dog 12 month

Chapter | 78  Dialkyldithiocarbamates (EBDCs)

1695

Table 78.2  (Continued) Study

Active

NOAEL (mg/kg/d)

LOAEL (mg/kg/d)

Effects observed at LOAEL/ critical results

Reference

Monkey 6 month

maneb

7.3

22

Increased thyroid weight

Leuschner et al. (1977)

metiram

5

15

Decreased T3, T4, increased thyroid weight and follicular hyperplasia

Sortwell et al. (1979)

mancozeb

17

170

Decreased weight gain

Shellenberger (1991)

13

130

Decreased weight gain, T3, T4

Everett et al. (1992)

maneb

11

44

Decreased body weight, T4; hepatocellular adenomas at 440 mg/kg body weight/day

Tompkins (1992)

metiram

24

79

Decreased body weight

Hunter et al. (1979)

mancozeb

4.8

29

Decreased weight gain, T3, T4; increased TSH, thyroid weight, follicular cell hypertrophy, hyperplasia, nodular hyperplasia, adenoma and carcinoma

Stadler (1990)

4

16

Decreased weight gain, T4; increased height of thyroid follicular epithelium, prominent microfollicles

Hooks et al. (1992)

maneb

20

67

Decreased body weight, T4; increased 131I half-life, thyroid weight

Leuschner et al. (1979, 1986a,b); Leuschner (1991)

metiram

3.1

12

Muscular atrophy

Hunter et al. (1981)

Mouse oncogenic

Rat chronic-oncogenic

NOAEL  no observed adverse effect level. LOAEL  lowest observed adverse effect level.

the primary thyroid condition, since anemia is a known manifestation of hypothyroidism in dogs (Duncan et al., 1994; Jain, 1986). The mechanistic linkage between the prolonged disruption of the hypothalamic–pituitary–thyroid (HPT) axis and thyroid neoplasia has been confirmed in studies of sulfamethazine (Hard, 1998). Indeed, thyroid gland neoplasia can be induced experimentally in laboratory animals simply by a a low iodine diet (Schaller and Stevenson, 1966), i.e., without exogenous agents, indicating that the ­neoplasia was induced by an internal factor. Supplementing the diet with thyroid hormone abolishes the ­neoplastic response indicating that the internal factor is TSH (Doniach, 1970). Thus, the sequence of events relating thyroid hormone inhibition via hormonal imbalance to the onset of pituitary and thyroid follicular neoplasia in rodents is well characterized, resulting in the inference that the threshold for the early steps in the sequence, particularly the key elevation of TSH levels, is necessarily a threshold for the remaining

steps in the process including carcinogenesis. For purposes of human oncogenic risk assessment, the principle of the existence of a threshold for thyroid and pituitary neoplasia resulting from thyroid inhibition has been accepted (EPA, 1998; Hard, 1998; Hill et al., 1989, 1998; IPCS, 1990), and the specific relevance of this threshold mechanism to ETU is also accepted (EPA, 1992, 1998; Hurley et al., 1998). In addition, in comparison to laboratory animals, humans are expected to exhibit a lesser degree of ­sensitivity to thyroid inhibitors (Costigan, 1998; Hard, 1998; Hill et al., 1998). The reasons for this are threefold: First, humans possess a substantial reserve supply of thyroid hormone, much of which is carried in the serum bound to thyroxine-binding globulin, a serum protein that is missing in laboratory rodents (Odell et al., 1967). Therefore, release of stored thyroid hormones maintains normal serum levels for weeks in euthyroid humans (Martindale, 1972) and for weeks to several months in hyperthyroid individuals (Odell et al., 1967), despite daily doses of antithyroid drugs sufficient to completely block synthesis. This ­protein

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1696

Table 78.3  ETU: Critical Findings in the Most Relevant Studies Type of study/species

NOAEL (mg/kg/d)

LOAEL (mg/kg/d)

Effects observed at LOAEL/ critical results

Reference

Mouse 90 day

1.7

17

Thyroid follicular hyperplasia and decreased colloid density in both sexes, and increased liver weights in females

O’Hara and DiDonato (1985)

Rat 90 day

1.7

8.5

Altered thyroid function and follicular hyperplasia

Freudenthal et al. (1977a)

Dog 90 day

0.39

6.0

Decreases in rbc parameters and increased cholesterol; thyroid and other effects at 80 mg/kg/d

Briffaux (1991)

Monkey 6 month

0.1–0.5

2.5

Increased iodine uptake, thyroid follicular hyperplasia and pituitary hypertrophy

Leber et al. (1978)

1 yr dog

0.18

1.8

Slightly reduced body weight gain, increased thyroid weights and hypertrophy with colloid retention, pigment accumulation in the liver

Briffaux (1992)

2 yr rat

0.25

1.25

Thyroid vacuolarity and hyperplasia

Graham et al. (1973, 1975)

2 yr rat

0.37

9.25

Thyroid, pituitary and liver effects, thyroid and pituitary tumors, decreased body weights in males

Schmid et al. (1992)

Mouse 2 year

17

17

Decreased T4, increased TSH and diffuse thyroid follicular cytoplasmic vacuolation at 17 mg/kg/day; tumors of the thyroid, liver and/or pituitary at 56 mg/kg/day and higher

Chhabra et al. (1992); NTP (1992)

Rat 2 year

1.1

1.1

Decreased T4, increased TSH and thyroid follicular hyperplasia at 1.1 mg/kg/day; thyroid tumors at 3.7 mg/kg/day and higher

Chhabra et al. (1992); NTP (1992)

0.11–0.43

1.1–4.3

Thyroid follicular hypertrophy and hyperplasia; no reproductive effect at 4.3–21 mg/kg/day, HDT

Dotti (1992)

Rat

5

10

Anomalies of the brain, neural tube and tail at 10 mg/kg/day; a higher frequency of delayed ossification of the parietal bone at 5 mg/kg/day judged consistent with or close to a NOAEL at this level; maternal NOAEL  40 mg/kg/day based on maternal lethality at 80 mg/kg/day

Khera 1973)

Rat

10

20

Dilation of the lateral ventricle

Teramoto et al. (1978)

Subchronic dietary

Chronic dietary

Oncogenicity

Reproductive Rat 2 generation

Developmental toxicity

Chapter | 78  Dialkyldithiocarbamates (EBDCs)

1697

Table 78.3  (Continued) Type of study/species

NOAEL (mg/kg/d)

LOAEL (mg/kg/d)

Effects observed at LOAEL/ critical results

Reference

Rat

5

10

Decreased fetal weights; hydrocephalus at 20 mg/kg/day. Maternal NOAEL  40 mg/kg/ day based on maternal deaths and reduced weight gain at 80 mg/kg/day

Chernoff et al. (1979)

Rat perinatal (exposure from day 7 gestation today 15 postpartum)

25

30

Hydrocephalus, failure to nurse, increased open field activity in males

Chernoff et al. (1979)

Rat

15

25

Dilation of brain ventricles; maternal NOAEL  35 mg/kg/day, HDT

Saillenfait et al. (1991)

Rabbit

40

80

Increased resorptions, decreased brain weights, degeneration of fetal kidney proximal tubules; maternal NOAEL  80 mg/kg/day, HDT

Khera (1973)

Hamster

90

270

Cleft palate, tail and skeletal anomalies, decreased fetal weight; maternal NOAEL  810 mg/kg/day, HDT

Teramoto et al. (1978)

Hamsters

100

NA

NOAEL

Chernoff et al. (1979)

Mice

100

200

Increased supernumerary ribs; maternal NOAEL  100 mg/kg/day based on increased relative liver weight

Chernoff et al. (1979)

Guinea pigs

100

NA

NOAEL

Chernoff et al. (1979)

Cat

120

NA

Fetal NOAEL; reported maternal toxicity inconsistent with any other reported studies of ETU

Khera and Iverson (1978)

lodine

Thyroid gland

T3, T4

Blood supply

T3, T4

Tissues and organs

TSH

Pituitary gland

TRH

Hypothalamus

T3, T4

T3 = Triiodothyronine (Thyroid hormone) T4 = Tetraiodothyronine (Thyroid hormone) TRH = Thyrotropin releasing hormone TSH = Thyroid stimulating hormone Figure 78.2  Thyroid-pituitary feedback mechanism.

is ­ missing in rodents, resulting in comparatively rapid hormone turnover, normally higher levels of TSH, and increased sensitivity to the effects of hormone depletion. Second, the molar concentrations of thiourea compounds required to inhibit thyroid peroxidase activity are far

smaller in rats than in monkeys or humans (Takayama et al., 1986), indicating that humans must be exposed at much higher levels to achieve the same degree of enzyme inhibition. Third, under conditions of prolonged thyroid insufficiency, caused for example by nutritional iodine deficiency,

1698

the primary human response is goiter rather than neoplasia (Hill et al., 1989, 1998). Despite extensive epidemiological studies, no convincing evidence has yet emerged to link iodine deficiency with human thyroid cancer (Hard, 1998). Thus, not only is a threshold model appropriate for hazard assessment of the effects of ETU, but also a large uncertainty factor is not needed to insure adequate protection of the human populations.

78.3.3.2  Liver Effects Although the level of ETU exposure resulting from bioconversion of the EBDCs at maximum tolerated doses is generally insufficient to produce tumors of the liver, ETU given directly does produce tumors of the liver in mice (Chhabra et al., 1992; Innes et al., 1969). Metabolism of the EBDCs to ETU is less extensive in mice (Cameron et al., 1990; Piccirillo et al., 1992), and induction of liver tumors with ETU has been observed only in mice and only at higher dietary levels that were also associated with centrilobular hepatocellular cytomegaly and increased functional demand (i.e., work-related stress to the liver), in addition to thyroid inhibition and sequelae. Hepatocellular tumors have not been seen in rats or hamsters. Although thyroid effects were present, no increase in the incidence of hepatocellular tumors was noted after two years of dietary feeding at 17–18 mg/kg bw/day (Chhabra et al., 1992). The precise mechanism of liver tumor formation with ETU in mice has not been fully elucidated. One hypothesis is that the liver tumors are related to stress on the HPT axis. The liver is subject to metabolic regulation by thyroid and pituitary hormones, and liver neoplasms have also been produced in mice by two other thionamides, 2-­thiouracil and 6-methyluracil, which also produce thyroid neoplasms in rats and mice (IARC, 1974). A second threshold hypothesis notes that ETU exhibits the hallmarks of phenobarbital-type liver promotion, including mixed function oxidase induction, sustained hepatomegaly, eosinophilic foci, and cellular proliferation (McClain, 1995; Whysner et al., 1996). Yet a third notes that redox cycling of ETU, with consumption of glutathione, has been observed with rat liver microsomes in vitro (Decker and Doerge, 1991), and that oxidative damage due to glutathione depletion is a commonly recognized threshold mechanism of tumor formation in animals. A number of additional factors also lead to the ­inference that these liver tumors are nonrelevant to human risk. First, there have been no reports associating the use of the structurally related propylthiouracil or other clinically administered thionamide drugs with an excess incidence of primary liver cancer in humans. Second, there is a wide variability in the incidence of liver tumors among various strains of mice, which is partly dependent on hormonal and/or nutritional factors, in addition to genetic factors. Genetic factors are particularly ­operative

Hayes’ Handbook of Pesticide Toxicology

in the case of B6C3F1 and other C3H-derived strains whose high and variable background tumor incidences indicate the presence of a significant population of “initiated” or latent tumor cells whose potential is readily expressed under stressed conditions of various origins. It has been acknowledged by numerous authorities that the induction of these tumors is of questionable or no relevance for assessment of oncogenic potential in human populations, where the background incidence of liver cancer is extremely low (e.g., IPCS, 1990; EU Directive 93/21/EEC; IARC, 1987). In conclusion, the available information on the pathobiology and mechanism of the ETU-induced liver tumors indicates they are nonrelevant for the assessment of human risk at doses below the threshold for conventional toxic responses in these organs.

78.3.3.3  Genotoxicity Few materials have been tested for mutagenic potential as exhaustively as the EBDCs and ETU. Among the more than 200 reported studies are more than sufficient numbers of qualified assays to ensure that each of the EBDCs and ETU individually and collectively have been adequately tested in a wide variety of in vitro and in vivo mutagenicity tests. Care must be taken in the qualification step to exclude flawed or otherwise deficient results. For example, because of the rapid degradation of EBDCs in dimethylsulfoxide (DMSO), and the accompanying rapid liberation of metal ions, EBDC studies in which DMSO has been used as a solvent are invalid. A weight of the evidence evaluation of the scientifically valid studies shows that, when properly tested in higher organism test systems, the EBDCs and ETU are not mutagenic in the two major endpoints used to assess genotoxicity, gene mutations and chromosomal damage, and further, they do not cause adverse effects in ancillary tests of genotoxic damage. Thus, the weight of the evidence indicates that the EBDCs and ETU are not mutagenic in mammalian systems (EBDC/ETU Task Force, 1992; Elia et al., 1995). This conclusion is shared by other organizations, for example, by the NTP (1992) which stated that “ethylenethiourea has been tested extensively for genotoxicity in a variety of in vitro and in vivo test systems, and the results, with few exceptions, are negative,” and the WHO (1994) which concluded that ETU was not genotoxic.

78.3.3.4  Bioequivalent Doses of the EBDCs   and ETU The bioconversion factors in rats and mice provide a quantitative basis for comparing the dose–responses of the EBDCs and ETU, and these comparisons provide quantitative support for the concept that the EBDCs’ toxic effects are directly related to their conversion to ETU. Example calculations for mancozeb, maneb, and metiram

Chapter | 78  Dialkyldithiocarbamates (EBDCs)

using the 7.5% bioconversion factor in rats and the 6.0% factor in mice are shown in Table 78.4. It is clear that the ETU-equivalent dose levels arising from the feeding of the EBDCs are similar enough to the corresponding LOAELs and NOAELs for ETU to justify the presumption of a cause and effect relationship.

78.3.4  Reproductive and Developmental Toxicity Reproductive outcome is generally unaffected by exposure to EBDCs or ETU. There were no effects on reproductive parameters, the microscopic appearance of the reproductive organs, or neonatal survival or growth resulting from exposure to any of the EBDCs or ETU at levels below those producing frank systemic toxicity in the adults (Tables 78.2 and 78.3). Further, unlike most pesticides, the potential for long-term effects of in utero exposure has been investigated. With the exception of a slight increase in thyroid tumor incidence in rats, in utero and perinatal exposure to ETU did not alter the incidences of tumors produced by postweaning lifetime exposures in either rats or mice (Chhabra et al., 1992). Developmental toxicity is observed as malformations and embryofetotoxic effects at maternally toxic dose levels with all three EBDCs in the rat (Table 78.2). The effects seen are qualitatively consistent with those produced by ETU, and the dose–response is consistent with their causation by bio-converted systemic ETU doses. Sensitivity varies with the species. Exposure to sufficient doses of ETU at the critical stages of pregnancy produces malformations in rats, predominantly those of the central nervous system and head. Related malformations are produced in hamsters although only at very high doses approaching maternally toxic levels. The mouse and rabbit are far less sensitive. Developmental effects, even at relatively high doses, are limited to findings indicative of embryofetotoxicity. The guinea pig and cat have not shown evidence of teratogenic or other developmental effects (Table 78.3; see also Ruddick and Khera, 1975; Khera, 1987). A relationship of the developmental effects to thyroid inhibition is indicated by several lines of evidence. The malformations produced by ETU exposure in vivo are those expected as the result of thyroid insufficiency. They occur only at doses in excess of those producing significant thyroid inhibition in adults, and they have been ­prevented, at least in part, by coadministration of thyroxine (Emmerling, 1978a, b). A key concern with thyroid inhibitors is that impaired thyroid function may alter hormone-mediated events ­during development, leading to permanent alterations in brain morphology and function (Cooper and Kavlock, 1997). The potential for this type of effect with ETU has been examined in a postnatal behavioral study in rats, resulting in a NOAEL of 25 mg/kg/day, fivefold higher than the 5 mg /kg/day

1699

Table 78.4  Bioequivalent Doses of the EBDCs and ETU Critical effect/ active

Critical doses (mg/kg/day) EBDC

ETU-equivalent to EBDCa

ETU

mancozeb

29

2.2

3.7

maneb

90

6.8

metiram

NA

NA

mancozeb

15

1.1

maneb

22.4

1.7

metiram

20

1.5

mancozeb

4.8

0.36

maneb

20

1.5

metiram

12

0.9

Rats Thyroid tumor LOAEL

Thyroid effect LOAEL 1.1

NOAEL 0.37

Mice Liver, thyroid tumor LOAEL mancozeb

NA

NA b

maneb

440

metiram

NA

NA

mancozeb

130

7.8

maneb

44

2.6

metiram

302

18

mancozeb

17

1.0

maneb

11

0.66

metiram

84

5.0

56c

26

Thyroid LOAEL 17

NOAEL 1.7

NA  not applicable. a EBDC mg/kg bw/day  ETU bioconversion factor of 7.5% in rats, 6% in mice. b Liver adenomas only. c Liver, thyroid, and pituitary tumors.

NOAEL for other types of developmental toxicity demonstrated in a companion study (Chernoff et al., 1979). Consistent with the findings with ETU, no developmental effects were observed with the EBDCs in studies in rabbits. Feed refusal, weight loss, and late-developing

1700

abortions observed in these studies are not relevant to human health risk. Being ruminants, rabbits are very sensitive to disruption of their gut microflora by antimicrobials and fungicides (ICH, 1994), and this response is well known to precipitate late term abortions (A. Hoberman, private communication).

78.3.5  Neurotoxicity The one feature of the EBDCs’ toxicological profile which is not explainable by the relationship to ETU is hindlimb paralysis and associated effects, including muscular atrophy with mancozeb and metiram and an effect on the retina with mancozeb. Hindlimb paralysis occurs at high doses with all of the EBDCs (LOAELs 60–340 mg/kg bw/day) and is a property of their common primary metabolite, ethylenebisisothiocyanate-sulfide (EBIS) (Freudenthal et al., 1977b), presumably related to the ability to release the carbon disulfide moiety (e.g., Johnson et al., 1998). Multiple exposures are generally required to produce the effect. Acute neurotoxicity testing of maneb produced no indication of an adverse effect at doses up to 2000 mg/kg (Nemec, 1993). No evidence of neurotoxicity has been observed with ETU, and a firm chronic NOAEL for neurotoxicity of 8.2 mg/kg/day has been demonstrated in perfusion neuropathology studies of mancozeb (Stadler, 1991).

78.3.6  Metabolites other than ETU In studies of metabolites other than ETU, EBIS further metabolized in rats and mice to ETU and ETU metabolites (Iverson et al., 1977; Jordan and Neal, 1979), and the thyroid functional changes which would be expected as a result of this metabolism were observed in addition to neurotoxicity (Freudenthal et al., 1977b). No evidence of tumorigenicity was observed with EU (Innes et al., 1969), and neither EU nor any of the other mancozeb and ETU metabolites tested exhibited any teratogenic potential (Ruddick et al., 1976b).

78.3.7  Hazard Characterization In summary, based on extensive data, the EBDCs do not pose a hazard of acute intoxication, genetic damage, or reproductive or developmental toxicity below levels that produce other kinds of toxicity in adults, or of significant systemic toxicity by the dermal route. There is no evidence of bioaccumulation. Repeated exposures to high doses of the EBDCs affect the thyroid, liver, and nervous systems in laboratory animals. The thyroid and liver effects are due to their metabolism in small amounts to ETU, which interferes with the synthesis of thyroid hormone and induces stressrelated liver growth. These effects are reversible when

Hayes’ Handbook of Pesticide Toxicology

exposures are brief or intermittent, but prolonged exposures can produce secondary changes, including anemia and thyroid, pituitary, and liver tumors in rodents. Available mechanistic information establishes a threshold for the thyroid and pituitary tumors and indicates that none of the tumor types are relevant for human risk assessment at likely exposure levels. Thus, neither the EBDCs nor ETU pose an oncogenic risk for humans. The EBDCs’ neurotoxic effects are shared with their common primary metabolite EBIS, although not with ETU, and a reliable NOAEL well above anticipated human exposure levels has been confirmed in specific neuropathology studies of mancozeb.

78.4  Dose–response 78.4.1  Noael and Acceptable Daily Intake—ETU The acceptable daily intake (ADI) or reference dose is defined as the approximate exposure which, if incurred daily over an entire lifetime, appears to be without appreciable risk for the general population, including all subgroups; i.e., it is the exposure level which provides a reasonable certainty of no harm and is therefore safe within the meaning of the U.S. Food Quality Protection Act and the WHO. When clinical studies in humans are inappropriate or unavailable, the ADI is estimated from reliably conducted toxicity studies in laboratory animals by taking the NOAEL associated with the most sensitive endpoint in the most sensitive species, and applying uncertainty factors to account for inter and intraspecies variability, and the need to protect the most sensitive individuals and subgroups, such as infants and children, among the general population. For ETU, inspection of the critical findings in the comprehensive laboratory animal studies summarized in Table 78.3 readily reveals that thyroid and related parameters are the most sensitive effects. The rat has generally been the most sensitive species to ETU followed closely by the dog and monkey, with the mouse relatively insensitive. An overall NOAEL of 0.4 mg/kg/day for the effects of ETU on the HPT axis, and therefore for the effects of ETU in laboratory animals, is supported by the 0.39 mg/kg/day NOAEL for hypothyroid changes including anemia in the 90 day and one year study of ETU in dogs (Briffaux, 1991,1992), the 0.37 mg/kg/day NOAEL for thyroid, ­pituitary, and liver effects in the two year chronic study in rats (Schmid et al., 1992), and the 0.11–0.43 mg/kg/day NOAEL for thyroid effects among the parents in the rat two-generation reproduction study (Dotti, 1992). Further, it is consistent with the weight of evidence of a comprehensive database, including the 1.1 mg/kg/day NOAEL for induction of thyroid tumors in the two year carcinogen bioassay in rats. The 1.1 mg/kg/day level was a LOAEL

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for thyroid hormonal changes and follicular hyperplasia (Chhabra et al., 1992). Taking into account a standard 10-fold safety or uncertainty factor for interspecies variability and a second 10-fold factor for intraspecies variation leads to the ADI of 0.004 mg/kg bw/day for ETU, as recommended by FAO/WHO (1994). Given the known lesser sensitivity of humans to thyroid effects, this is a very conservative assessment, affording extra protection of the public health. A number of factors, among them the unusually complete and reliable database, including chronic effects of in utero exposure, the lesser sensitivity of humans to thyroid effects, and the fact that the NOAELs for developmental and reproductive effects are 10-fold or more higher than those for thyroid effects, confirms that no additional uncertainty factor is needed to ensure adequate protection of infants and children.

78.4.2  Noael and Acceptable Daily Intake—EBDCS The key requirements to support the identification of an aggregate NOAEL for a group of compounds are a common toxicological profile and a similar dose–response. As noted above, the EBDCs share a comparable toxicological profile, primarily based on the toxic effects of their common metabolite ETU. An issue which deserves some comment in this context is muscular atrophy, which was the critical

1.0 mg/kg bw/day

effect in the two year study of metiram in rats. Muscular atrophy has been seen at high doses with both mancozeb and metiram and is associated in both of these materials with clinically diagnosed hindlimb paralysis at the same or higher doses. Thus, this effect is not an exception to the common toxicological profile in the qualitative sense, although its appearance at the LOAEL for metiram was not typical of the dose–response seen with the other EBDCs. With this one possible exception, the three EBDCs exhibit similar dose–effect and dose–response curves, as is illustrated in Fig. 78.3, which presents the NOAELs and LOAELs for the critical studies from Table 78.2 in a graphical format. Inspection of the graph readily reveals that the dose–responses of the three EBDCs are consistent. With only minor exceptions, the NOAELs and LOAELs of the respective materials do not overlap for any particular type of study. Second, no one of the EBDCs is consistently more sensitive than the others. Of the 10 studies ­conducted with at least two of the three EBDCs, the LOAELs for each study type are evenly divided among the three (4 ­ mancozeb, 3 maneb, 3 metiram), indicating that the apparent differences among the compounds in individual studies are due to dose selection, and not to any intrinsic differences in potency. Similarly, the critical LOAELs for all three compounds are nearly identical, at between 10 and 16 mg/kg bw/day. Since the toxicological profiles and dose–response are similar for the three EBDCs, it is appropriate to base the determination of the overall NOAEL for the EBDCs on

10.0 mg/kg bw/day

100.0 mg/kg bw/day

Mouse 3 month Rat 3 month Dog 3 month Dog 12 month Monkey 6 month Mouse oncogenic Rat chroniconcogenic Rat 2 generation Rat neuropathology Rat teratology Rabbit teratology Overall NOAEL for EBDCS = 5.0 mg/kg bw/day

Figure 78.3  Summary of critical doses in toxicology studies of EBDCs.

Mancozeb NOAEL

Mancozeb LOAEL

Maneb NOAEL

Maneb LOAEL

Metiram NOAEL

Metiram LOAEL

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the aggregated dataset for the group. An overall NOAEL of 5.0 mg/kg/day for the EBDCs as an aggregate group is supported by all of the available studies, as presented in Table 78.2 and Fig. 78.2. It is further supported by consideration of the dose response and NOAELs for ETU. Taking into account the 7.5% bioconversion factor (EPA, 1992; Kocialski, 1989), a 5.0 mg/kg/day dose of EBDC would result in a systemic dose of 0.4 mg/kg/day of ETU (5.0 mg/kg/day  0.075  0.375 mg/kg/day), precisely equal to the overall NOAEL for ETU determined in independent studies. This comparison lends the support of the full ETU database to the 5.0 mg/kg/day aggregate NOEL for the EBDCs. In summary, an aggregate evaluation of the individual toxicology databases for the respective EBDCs, mancozeb, maneb, and metiram, and their ETU metabolite indicates an overall NOAEL of 5.0 mg/kg bw/day for the group. Application of a standard 100-fold overall safety factor leads to a recommended acceptable daily intake of 0.05 mg/kg/day. The factors discussed above for ETU, and the unusual reliability of the database, reflecting the combined results of four individually comprehensive databases, confirm that no additional uncertainty factors need be applied. This aggregate assessment produces a comparable, if slightly higher, estimate of the ADI than the FAO/WHO recommendation. The WHO panel reviewed the data for each of the actives individually and established ADIs of 0.05 mg/kg/day for mancozeb and maneb, and 0.03 mg/kg/day for metiram, resulting in the allocation of a group ADI of 0.03 mg/kg/day for the EBDCs collectively, including mancozeb, maneb, metiram, and zineb. The basis for the establishment of a group ADI was the similarity of the chemical structures of the EBDCs, the comparable toxicological profile of the EBDCs based on the toxic effects of ETU, and the fact that the parent EBDC residues cannot be differentiated using presently available regulatory analytical procedures (FAO/WHO, 1994).

78.4.3  Acute Reference Dose The acute reference dose (aRfD) is the maximum single day oral exposure which is anticipated to be without appreciable risk for the general population. Toxic effects which might occur as a result of exposures occurring within the period of a single day, or after at most a very few doses, are relevant to the assessment. For ETU the only relevant acute toxicological endpoint is developmental toxicity. A NOAEL of 5.0 mg/kg/day is supported by the weight of the evidence of multiple ­studies in the rat, as the most sensitive species (Chernoff et al., 1979; Khera, 1973; Saillenfait et al., 1991; Teramoto et al., 1978). The mechanistic relationship to thyroid inhibition suggests that multiple exposures, producing hormone depletion, would be required for full expression of ETU’s

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developmental toxicity potential in the lower dose range. This and the unusually thorough nature of the database argue that the standard 100-fold uncertainty factor is more than adequate to assure protection of women, infants, and children, indicating an aRfD for ETU of 0.05 mg/kg/day. For the EBDCs, since expression of their neurotoxic potential requires repeated doses (e.g., Nemec, 1993) and their relevant developmental effects are due to bioconversion to ETU, the most relevant endpoint for assessment of acute risks is the aRfD for ETU of 0.05 mg/kg bw/day, applied to the combined direct and indirect (7.5% bioconverted EBDC dose) ETU exposure.

78.4.4  Endpoints for Assessment of Dermal and Respiratory Exposure For assessment of the potential risks of pesticide users and bystanders, and those who encounter exposure after ­application, the dermal and respiratory routes of exposure are most relevant. Dermatitis due to repeated exposures and irritation of the mucus membranes are prevented by appropriate personal protective equipment. Dermal absorption of the EBDCs is low, 6.5% or less, and consistently with this, the EBDCs are of very low toxicity by the dermal route, with NOAELs in 21-day dermal toxicity studies ranging from 100 mg/kg/day for maneb to 1000 mg/kg/day (limit dose) for mancozeb. Alternatively, the aRfD and ADIs derived from the oral exposure studies may be applied to the estimated systemic exposures from single or multiple doses, respectively, after adjustment for dermal absorption. Subchronic inhalation studies produced the same kinds of effects as in oral studies of the same length and a generally similar dose–response, with NOAELs of 79, 10, and 2 mg/ m3 for mancozeb, maneb, and metiram, respectively. After correction for respiration rates and respirable fraction, the estimated systemic exposure of 8.3 mg/kg/day at the NOAEL in the mancozeb study was equal to the 7.4–8.2 mg/kg/d NOAELs in the 90-day dietary studies, indicating the relevance of the aRfD and ADIs for systemic exposures. In enclosed spaces, a Workplace Environmental Exposure Limit Guide of 1 mg/m3 is recommended (AIHA, 1992).

78.4.5  Carcinogenicity Classification and Low Dose Risk Assessment The criteria for classification of substances as carcinogens differ with the classifying authority. Mancozeb and ETU were evaluated by the EU Commission in 1994, in the context of European classification criteria which assign an important role to the mode of action, and were not classified as carcinogens. In contrast, the International Agency for Research on Cancer (IARC) classified ETU as category 2B “possibly carcinogenic to humans” based on sufficient evidence

Chapter | 78  Dialkyldithiocarbamates (EBDCs)

in laboratory animals but inadequate evidence in humans (IARC, 1987), and the EPA (1992) cited evidence of carcinogenicity in two species and a weak genotoxic potential in classifying ETU, as a B2 “probable human carcinogen.” The Agency’s classification of the EBDCs as similar B2 carcinogens was largely due to the metabolic conversion to ETU. Because EPA prefers a probabilistic approach to risk assessment for carcinogens in this class, the linear multistage model was used to calculate an upper 95% confidence limit of 0.06 (mg/kg bw/day)1 on the lifetime risk of cancer from ETU at low exposure levels (q*) (EPA, 1997).

78.5  Toxicology in humans With the exception of sporadic reports of allergic contact hypersensitivity (Bruze and Fregert, 1983; Crippa et al., 1990; Kleibl and Rackova, 1980; Lee et al., 1981), studies of manufacturing workers and users have not discerned adverse effects of exposure to either ETU or the EBDCs. In the most comprehensive study of thyroid effects, 153 men currently or previously exposed to EBDC for many years at a manufacturing site were compared for thyroid function to 153 men not exposed to EBDC, its products, or ETU who also worked at the same plant. Workers and controls were carefully matched with respect to age, race, length of employment, and type of employment. Informed consent was obtained from all participants. No significant differences were observed on thyroid palpitation and in thyroid function tests between the EBDC workers and the control men. Urinary excretion of ETU in a subgroup of 42 workers currently exposed to EBDC was 0.002 ppm compared to 0.001 ppm in the control group (41 men), most of whom had undetectable values. The authors concluded that exposure to EBDC manufacture was not associated with an increased prevalence of thyroid abnormalities (Charkes et al., 1985). Clinical examinations and thyroid function tests were also carried out over a period of 3 years in the United Kingdom on 8 male workers engaged for between 5 and 20 years in the manufacture of ETU and 5 male workers involved for 3 years in the mixing of ETU with rubber, and matched controls. Levels of ETU recorded on personal samplers of manufacturing workers reached 330 (g/m3, and levels for mixers ranged from 120 to 160 g/m3. Mixers but not process workers had significantly lower levels of T4 in their blood compared to controls. With the exception of one mixer with elevated TSH levels, who was evaluated as hypothyroid on further testing, no effects were found on TSH or thyroid binding globulin. The authors concluded that there was no evidence that thyroid function is severely affected by exposure to ETU at the levels experienced by these workers (Smith, 1984). Similarly, no hazard of clinical thyroid depression existed based on medical evidence

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collected on 51 workers exposed to ETU at a U.S. rubberprocessing company (Salisbury and Lybarger, 1977). No difference was found in the total death rate or deaths due to cancers between 992 male workers involved in the production of EBDCs from 1948 to 1975 and control males from the city of Philadelphia. No thyroid cancer was found in this study (DeFonso, 1976). Epidemiological studies were conducted on workers in the rubber industry by Parkes (1974) and Smith (1976). Based on examination of national and regional thyroid cancer incidences in the UK, Parkes concluded that, under the conditions in which ETU had been used in the rubber industry, there is no risk of man contracting thyroid cancer as a result of industrial exposure to ethylenethiourea (Parkes, 1974). Similarly, a total of 1929 workers engaged in the production or manufacture of ETU were surveyed retrospectively for thyroid cancer and were compared with the thyroid cancer list of the Birmingham (England) Cancer Registry from 1957 to 1971. No thyroid cancers occurred in these workers. A retrospective study of 699 women who were employed at a rubber manufacturer using ETU assessed the incidence of fetal abnormalities occurring in children to women who had worked with ETU during early pregnancy. No excess incidence was found, and the study did not demonstrate any risk of teratogenesis (Smith, 1976).

78.6  Risk characterization 78.6.1  Dietary Exposure and Risks Risk is a function of hazard potential and exposure. The potential risk of consumers of foods derived from chemically protected crops is assessed by comparing the ADI for the particular crop protection chemical, or other indices of its toxicity potential when relevant, to estimated or measured dietary exposure values. Dietary exposure is in turn determined from the level of residues of the crop protection chemical and its toxicologically significant metabolites in foods and food consumption patterns for various population groups. Dietary intake may be predicted with varying degrees of accuracy. The WHO and the European Union use the International Estimated Dietary Intake (IEDI) for five global diets. The exposure is calculated using ­ average regional food consumption data and the median residue from supervised trials that have been used to assess the maximum residue limit or tolerance consistent with the worldwide labels. Residues of EBDCs and ETU in ­processed commodities were also calculated. The IEDI was calculated for each of the five global diets using the EBDC and ETU level with the highest dietary contribution. On this basis, the aggregate dietary exposure to the EBDCs ranged from a low of 3% of the ADI in the African diet to 36% of the ADI in the European diet. These levels are well

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within the criteria of WHO for fully acceptable dietary risk (less than 100% of the ADI) (Ollinger, 1999). These data overestimate the dietary risks of the EBDCs because the IEDI calculation assumes that 100% of the crop is treated and the residues are measured from crops as they are harvested. More reliable assessments of dietary exposure may be obtained at the national level (National Estimate of Daily Intake, NEDI) . Such a more accurate determination of EBDC and ETU exposure is available from a market basket survey conducted in the United States. In that study, samples of fresh and processed foods were collected every two weeks for one year from grocery stores throughout the United States in a statistically designed survey. Using the most sensitive residue analytical methods achievable, the results from the analysis of almost 6000 samples showed that about 80% of the samples had no measurable residues of EBDC or ETU. The residues that were found were very close to the limit of quantitation in the method. Thus, there is virtually no exposure to the consumer to residues of either EBDCs or ETU. A reliable NEDI was derived for the United States, a member of the European diet category, from the average consumer level residue values determined in the market basket and related studies, and detailed national data on the consumption of raw and processed food commodities as summarized in the US EPA’s Dietary Risk Estimation System. The NEDI for U.S. consumer exposure to EBDCs and ETU was 0.000027 mg/kg bw/day, calculated as ETU, or less than 1% (0.68%) of the 0.004 mg/kg bw/day ADI for ETU. These NEDI values clearly show that actual consumer exposure to residues of the EBDCs is negligible (EBDC/ETU Task Force, 1997b). Utilization of these market basket data in a Monte Carlo assessment of acute dietary exposure resulted in an acute margin of exposure for women of child-bearing age (13  years) of 2652 at the 99.9 percentile of exposure, greatly beneath any levels of concern (EBDC/ETU Task Force, 1997b). EPA calculated the risks to consumers using the market basket data and their standard method of estimating the probability of human cancer risks. They determined that there was only a negligible theoretical dietary risk of 1.2 in one million from the use of EBDCs on 48 crops (EPA, 1992, 1996). The negligible character of dietary exposure to the EBDCs is further underlined by the nutritional benefits which they confer. The principal use of the EBDCs is the economical protection of fruit and vegetables from disease, making these commodities more widely available in the diet. Overwhelming evidence from epidemiological studies indicates that diets high in fruits and vegetables are associated with a lower risk of numerous cancers, and for this reason, dietary recommendations to increase the intake of citrus fruits, cruciferous vegetables, green and yellow

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v­ egetables, and fruits and vegetables high in vitamins A and C have been made by numerous organizations (AICR, 1997; American Cancer Society, 1984, 1996; Block et al., 1992; Giovannucci, 1999; NAS/NRC, 1989; NCI, 1987; Steinmetz and Potter, 1991).

78.6.2  Worker Exposure and Risks Exposure estimates for mixing, loading, and applying EBDC fungicides during outdoor agricultural applications have been calculated using the predictive operator exposure model of the UK MAFF and also the model of the German BBA as sources of surrogate exposure data. Each of these models represents a synthesis and integration of exposure data obtained in a large number of field trials conducted using various kinds of equipment in a variety of national agricultural settings. The final models were calibrated to insure that the resulting estimates would, if anything, overpredict actual exposures under field conditions. For modeling of EBDC uses, an application rate of 2.4 kg ai/hectare was chosen to represent a typical maximum use rate of EBDC products when used on a standalone basis. Use rates for these products when used as mixtures are lower. As gloves are recommended as personal protective equipment when handling undiluted product, the use of gloves (during mixing and loading only) is presumed in the models. The exposure estimates obtained using these very conservative models indicate that operator exposures are always below, and in most cases substantially below, acceptable dermal and inhalation exposure limits for workers, even when personal protective equipment is limited to the bare minimum of gloves during mixing and loading. Therefore a significant risk for the operator during the use of EBDC fungicide products appears unlikely. As in the case of dietary exposure, alternative risk assessment approaches have been used at times by various authorities. At the conclusion of the U.S. Special Review, the EPA concluded there was adequate safety to mixers, loaders, and applicators when proper personal protective equipment is used (EPA, 1992). Estimates of exposure for other activities, including bystanders, workers reentering treated fields, homeowners, and others reentering treated areas are even lower.

CONCLUSION Because of their importance to worldwide agriculture, the EBDCs and ETU have been thoroughly tested over many years. Collectively, the data demonstrate that the use of the EBDCs results in essentially negligible exposure to consumers, coupled with a significant contribution to improved nutrition, and low risk to farm workers, production workers, and people who are exposed through recreational activities.

Chapter | 78  Dialkyldithiocarbamates (EBDCs)

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Innes, J. R. M., Ulland, B. M., Valerio, M. G., Petrucelli, L., Fishbein, L., Hart, E. R., Pallotta, A. J., Bates, R. R., Falk, H. L., Gait, J. J., Klein, M., Mitchell, I., and Peters, J. (1969). Bioassay of pesticides and industrial chemicals for tumorigenicity in mice; a preliminary note. J. National Cancer lnst. 42, 1101–1114. International Agency for Research on Cancer (IARC) (1974). “Ethylenethiourea,” Vol. 7, pp. 23–24 and 45–52. Lyon, France. International Agency for Research on Cancer (IARC) (1987). “Monographs on the Evaluation of Carcinogenic Risks to Humans. Genetic and Related Effects: An Updating of Selected IARC Monographs from Volumes 1–42,” Suppl. 6, pp. 304–307. Lyon, France. International Conference on Harmonization (ICH) (1994). “Guideline on Detection of Toxicity to Reproduction for Medicinal Products.” United States Food and Drug Administration, 59 Federal Register 48746, September 22, 1994. International Programme on Chemical Safety (IPCS) (1990). . In Environmental Health Criteria 104. Principles for the Toxicological Assessment of Pesticide Residues in Food, pp. 88–91. World Health Organization, Geneva, Switzerland. Iverson, F., Newsome, W. H., and Hierlihy, S. L. (1977). Tissue distribution of ethylenefhiuram monosulfide (ETM) in the rat. Bull. Environmental Contarn. Toxicol. 18, 541–551. Iverson, F., Khera, K. S., and Hierlihy, S. L. (1980). In vivo and in vitro metabolism of ethylenethiourea in the rat and the cat. Toxicol. Appl. Pharmacol. 52, 16–21. Jackh, R. (1981). “Report on the Study of the Acute Oral Toxicity of Metiram Technical Grade in the Rat (Translation).” Unpublished Rep. BASF 92/10669 of BASF Corporation. Jain, N. C. (1986). In Schalm’s Veterinary Hematology, 4th ed, p. 671. Lea and Febiger, Philadelphia. Johnson, D. J., Graham, D. G., Amarnath, V., Amarnath, K., and Valentine, W. M. (1998). Release of carbon disulfide is a contributing mechanism in the axonopafhy produced by N,N-diethyldifhiocarbamate. Toxicol. Appl. Pharmacol. 148, 288–296. Jordan, L. W. and Neal, R. A. (1979). Examination of the in vivo metabolism of maneb and zineb to ethylenethiourea (ETU) in mice. Bull. Environmental Contam. Toxicol. 22, 271–277. Kapp, R. W., Schellhaas, L. J., and Piccirillo, V. J. (1991). “Prenatal Toxicity Study of Maneb in Rats.” Unpublished Reps. BASF 88/0522, 88/0523, 88/0524 of BASF Corporation. Kato, Y., Odanaka, Y., Teramoto, S., and Matano, O. (1976). Metabolic fate of ethylenethiourea in pregnant rats. Bull. Environmental Contam. Toxicol. 16, 546–555. Khera, K. S. (1973). Ethylenethiourea: Teratogenicity study in rats and rabbits. Teratology 7, 243–252. Khera, K. (1987). Ethylenethiourea: A review of teratogenicity and distribution studies and as assessment of reproduction risk. CRC Crit. Rev. Toxicol. 18, 129–139. Khera, K. S. and Iverson, F. (1978). Toxicity of ethylenethiourea in pregnant cats. Teratology 18, 311–314. Kleibl, K., and Rackova, M. (1980). Cutaneous allergic reactions to dithiocar-bamates. Contact Dermatitis 6, 348–349. Klimisch, H. J., and Zeller, H. (1980). “Report on the Determination of the Acute Inhalation Toxicity LC50 Metiram Technical with 2% ETU as a Dust Aerosol after 4-hr Exposure in Sprague-Dawley Rats.” Unpublished Rep. RZ 83/064 of BASF Corporation. Kocialski, A. (1989). “Establishment of an In Vitro Metabolic Conversion Factor of 7.5% for all EBDCs when Converting EBDCs to Ethylenethiourea In Vivo and Recalculations of the Previously

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Considered 20% In Vivo Conversion/Exposure Factor for EBDCs to ETU,”. Memorandum to Caswell file for each EBDC and ETU, Health Effects Division, Office of Pesticide Programs, United States Environmental Protection Agency. Leber, A. P., Wilkinson, G. E., Persing, R. L., and Holzworth, D. A. (1978). “Effects of Feeding Ethylene Thiourea in the Rhesus Monkey.” Contract 68-01-4717 from Battelle Laboratories, Columbus, OH. Unpublished report of Rohm and Haas Company. Lee, Y. S., Cinn, YW., Chang, W. H., and Kim, J. S. (1981). A study on hyper-sensitivity of Korean farmers to various agrochemicals. 1. Determination of concentration for Patch tests of fruit-tree agrochemicals and hypersen-sitivity of orange orchard farmers in Che-Ju Do, Korea.. Seoul J. Med. 22, 137–142. Leuschner, F. (1979a). “The Acute Oral Toxicity of the Preparation Metiram, Technical Agent with 2% ETU in Rats.” Unpublished Rep. BASF 79/0161, WNT-Nr. 77/951 of BASF Corporation. Leuschner, F. (1979b). “The Acute Intraperitoneal Toxicity of the Preparation Metiram, Technical Agent with 2% ETU in Mice.” Unpublished Rep. BASF 79/0162, WNT-Nr. 77/951 of BASF Corporation. Leuschner, F. (1991). “Chronic Oral Toxicity of Manganese-ethylene-l,2-bis-dithiocarbamate, 90%—Called for Short ‘Maneb’— Sprague-Dawley (SIV) Rats (with Special Attention to Carcinogenic Properties). Amendment 3.” Unpublished registration document (BASF) 86/0430 of BASF Corporation. Leuschner, F, Leuschner, A., Schneider, C, Schwerdtfeger, W., and Dontenwill, W. (1977). “Oral Toxicity of Maneb in the Rhesus Monkey: Six Months Dietary Dosing.” Unpublished Rep. WF 1172 of BASF Corporation. Leuschner, F, Leuschner, A., Klie, R., Dontenwill, W., and Rogulja, P. (1979). “Chronic Oral Toxicity of Manganese-ethylene-1,2-bis-dithiocarbamate, 90%—Called for Short ‘Maneb’—Sprague-Dawley (SIV) Rats (with Special Attention to Carcinogenic Properties).” Unpublished registration document (BASF) 86/0430 of BASF Corporation. Leuschner, E, Dontenwill, W., and Hubschner, F. (1986a). “Chronic Oral Toxicity of Manganese-ethylene-1,2-bis-dithiocarbamate, 90%— Called for Short ‘Maneb’—Sprague-Dawley (SIV) Rats (with Special Attention to Carcinogenic Properties). Amendment 1.” Unpublished registration document (BASF) 86/0430 of BASF Corporation. Leuschner, F., Dontenwill, W., and Hubschner, F. (1986b). “Chronic Oral Toxicity of Manganese-ethylene-1,2-bis-dithiocarbamate, 90%— Called for Short ‘Maneb’—Sprague-Dawley (SIV) Rats (with Special Attention to Carcinogenic Properties). Amendment 2.” Unpublished registration document (BASF) 86/0430 of BASF Corporation. Lewerenz, H. J., and Plass, R. (1984). Contrasting effects of ethylenethiourea on hepatic monooxygenase in rats and mice. Arch. Toxicol. 56, 92–95. Longacre, S. (1986). “Summary of ETU and EBDC Analysis in Plasma, Liver, and Thyroid after Mancozeb Administration. Supplement/appendix to Nelson (1986). Rep. 31H-86-02. Unpublished Rep. 86R-1009 of Rohm and Haas Company. Martindale, W. B. (1972). Carbimazole and other antithyroid agents. In “Extra Pharmacopoeia,” (N. W. Blacow ed.), 26th ed, pp. 379–385. Pharmaceutical Press, London. Matsushita, T., Arimatsu, Y., and Nomura, S. (1976). Experimental study on contact dermatitis caused by dithiocarbamates maneb, mancozeb, zineb, and their related compounds. Internal Arch. Occupational Environmental Health 37, 169–178. Matsushita, T., Yoshioka, M., Arimatsu, Y., and Nomura, S. (1977). Experimental study on cross-contact allergy due to dithiocarbamate fungicides. Industrial Health 15, 87–94.

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McClain, R. M. (1995). Phenobarbital mouse liver tumors: Implications of hepatic tumor promotion for cancer risk assessment. In Growth Factors and Tumor Promotion: Implications for Risk Assessment, pp. 325–336. Wiley-Liss, New York. Merkle, J. (1983). “Study to Determine the Prenatal Toxicity of Manganous Ethylenebis (Dithiocarbamate) in Rabbits.” Study RZNo. 83/094 from BASF AG, Ludwigshafen Rhein, Federal Republic of Germany. Unpublished report of Elf Atochem North America, Inc. Muller, W. (1991). “Mancozeb Oral (Gavage) Teratogenicity Study in the Rabbit.” Rep. HLD-853-683-002 from Hazleton Labs, Deutschland Gmbh, Munster, Germany. Unpublished report of Elf Atochem North America, Inc. Muller, W. (1992). “Mancozeb 2-Generation Oral Reproduction Toxicity Study in the Rat.” Rep. HLD-852-683-001 from Hazleton Labs, Deutschland Gmbh, Munster, Germany. Unpublished report of Elf Atochem North America, Inc. Naas, D. J. (1989a). “Acute Oral Toxicity (LD50) Study in Albino Rats with Maneb Technical.” Rep. WIL-134003 from WIL Research Laboratories, Inc., Ashland, OH. Unpublished report of Elf Atochem North America, Inc. Naas, D. J. (1989b). “Acute Dermal Toxicity (LD50) Study in Albino Rabbits with Maneb Technical.” Rep. WIL-134004 from WIL Research Laboratories, Inc., Ashland, OH. Unpublished report of Elf Atochem North America, Inc. National Academy of Sciences, National Research Council(1989). “Diet and Health: Implications for Reducing Chronic Disease Risk,” National Academy Press, Washington, DC. National Cancer Institute (NCI)(1989). “Diet, Nutrition, and Cancer Prevention: A Guide to Food Choices,” United States Government Printing Office, Washington, DC. National Toxicology Program (NTP) (1992). “Toxicology and Carcinogenesis Studies of Ethylene Thiourea in F344/N Rats and B6C3F1 Mice (Feed Studies).” United States Department of Health and Human Services, Public Health Service, National Institutes of Health Publication 92-2843, United States National Toxicology Program Technical Rep. Sen 388. Nelson, S. S. (1986). “Metabolism of 14C Mancozeb in Rats.” Unpublished Rep. 31H-86-02 of Rohm and Haas Company. Nelson, S. S. (1987). “Bioconversion of Mancozeb to ETU in Rat.” Unpublished Rep. 31C-87-24 of Rohm and Haas Company. Nemec, M. D. (1992). “A Developmental Toxicity Study of Maneb Technical in Rats.” Rep. WIL-134011 from WIL Research Laboratories, Inc., Ashland, OH. Unpublished report of Elf Atochem North America, Inc. Nemec, M. D. (1993). “An Acute Neurotoxicity Study of Maneb Technical in Rats.” Rep. WIL-134015 from WIL Research Laboratories, Ashland, OH. Unpublished report of Elf Atochem North America, Inc. Odell, W. D. et al. (1967). Studies of thyrotropin physiology by means of ra-dioassays. In “Proceedings of the 1986 Laurentian Hormone Conference (G. Pincus ed.)Vol. 23, pp. 47–85. Academic Press, San Diego. O’Hara, G. P., and DiDonato, L. J. (1985). “Dithane M-45 and Ethylenethiourea: 3-Month Dietary Toxicity Study in Mice.” Unpublished Rep. 80R-124 of Rohm and Haas Company. Ollinger, J. (1999). “Dithiocarbamates (CCPR Code 105). EBDC/ETU, STMR-P for Apple Juice, EBDC IEDI Calculation.” Letter to W. H. van Eck, Codex Committee on Pesticide Residues, EBDC/ETU Task Force. O’Neil, W., and Marshall, W. (1984). Goitrogenic effects of ethylenethiourea on rat thyroid. Pesticide Biochem. Physiol. 21, 92–101.

Chapter | 78  Dialkyldithiocarbamates (EBDCs)

Palmer, A., and Simons, R. (1979). “Effect of Metiram Technical on Pregnancy of the Rat.” Rep. RZ 79/065, BSF 302/79616 from Huntingdon Research Centre, Huntingdon England. Unpublished report of BASF Corporation. Parkes, H. G. (1974). Living with carcinogens. J. Inst. Rubber Industry 8, 21–23. Peters, A. C, Kurtz, P. J., Donorrio, D. J., Thake, D. C, and Cottrill, D. L. (1980a). “Prechronic Studies of Ethylenethiourea: Acute, Repeated Dose and Subchronic in Rats.” Unpublished Project G-7186 of Battelle Laboratories, Columbus, OH, for United States National Institute of Environmental Health Sciences. Peters, A. C, Kurtz, P. J., Donorrio, D. J., Thake, D. C, and Cottrill, D. L. (1980b). “Perchronic Studies of Ethylene Thiourea: Acute, Repeated Dose and Subchronic in Mice.” Unpublished Project G-7186 of Battelle Laboratories, Columbus, OH, for United States National Institute of Environmental Health Sciences. Peters, A. C, Kurtz, P. J., Chin, A. E., Carlton, B. D., Chrisp, C. E., and Dill, G. S. (1982). “Report on the Maximum Neonatal Dose Studies with Ethylenethiourea.” Unpublished Contract N01-ES-8-2151 of Battelle Laboratories, Columbus, OH, for United States National Institute of Environmental Health Sciences. Piccirillo, V. J., Wu, D., and Speirs, G. (1992). “Metabolism of Mancozeb in the Mouse.” Rep. T91-3413 from Inveresk Research International, Ltd. Tranent, Scotland. Unpublished report of Elf Atochem North America, Inc. Puhl, J. R. (1985). “Metabolism of Radiolabelled Maneb in Rats.” Rep. 6181-101 from Hazleton Laboratories, America, Inc., Madison, WI. Unpublished report of Elf Atochem North America, Inc. Rowland, J.(1997). “Toxicity Endpoint Selection Process,” Health Effects Division, United States Environmental Protection Agency Office of Pesticide Programs. Ruddick, J. A., and Khera, K. S. (1975). Pattern of anomalies following single oral doses of ethylenethiourea to pregnant rats. Teratology 12, 277–281. Ruddick, J. A., Williams, D. T., Hierlihy, L., and Khera, K. S. (1976a). 14C-ethylenenthiourea: Distribution, excretion, and metabolism in pregnant rats. Teratology 13, 35–40. Ruddick, J. A., Newsome, W. H., and Nash, L. (1976b). Correlation of teratogenicity and molecular structure: ethylenethiourea and related compounds. Teratology 13, 263–266. Ruddick, J. A., Newsome, W. H., and Iverson, F. (1977). A comparison of the distribution, metabolism and excretion of ethylenethiourea in the pregnant mouse and rat. Teratology 16, 159–162. Ryle, P. R., Bell, P. F., Parker, C, Farmer, H., Offer, J. M., Anderson, A., and Dawe, I. S. (1991). “A Study of the Effect of Maneb (Technical) on Reproductive Function of Two Generations in the Rat.” Rep. MNB1/9072 from Huntington Research Centre Ltd., Cambridgeshire, England. Unpublished report of Elf Atochem North America, Inc. Saillenfait, A. M., Sabate, J. P., Langonne, I., and De Ceaurriz, J. (1991). Difference in the developmental toxicity of ethylenethiourea and three N,N-substituted thiourea derivatives in rats. Fundamental Appl. Toxicol. 17, 399–408. Salisbury, S. A., and Lybarger, J. (1977). “Ethylene Thiourea.” Health Hazard Evaluation Determination Rep. 77-67-499 from St. Clair Rubber Company, Marysville, MI, United States Department of Health, Education and Welfare. Schaller, R. T., and Stevenson, J. K. (1966). Development of carcinoma of the thyroid in iodine-deficient mice. Cancer. 19, 1063–1080. Schmid, H., Tennekes, H., Janiak, T., Probst, D., Luefkemeier, H., Pappritz, G., Marki, U., Vogel, O., and Heusner, W. (1992).

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“Ethylene Thiourea 104 Week Chronic Toxicity (Feeding) Study in Rats.” Project 256803 from Research and Consulting Company, Ltd., Itingen, Switzerland. Unpublished report of ETU Task Force. Shaw, D. (1990). “Mancozeb: 52-Week Oral (Dietary) Study in the Beagle.” Project HLA 5913-616/3 from Hazleton Laboratories, UK. Unpublished report of Rohm and Haas Company. Shellenberger, T. (1991). “Mancozeb: 18-Month Oncogenicity Study in Mice.” Project 85051 from Tegeris Labs, Inc., Temple Hills, MD. Unpublished report of Rohm and Haas Company. Smith, D. (1976). Ethylene thiourea—A study of possible teratogenicity and thyroid carcinogenicity. J. Soc. Occupational Med. 26, 92–94. Smith, D. (1984). Ethylene thiourea: Thyroid function in two groups of exposed workers. British J. Industrial Med. 41, 362–366. Solomon, H., and Holz, J. (1987). “Mancozeb: Range-Finding Developmental Study in Rats.” Unpublished Rep. 85R-244 from Rohm and Haas Company. Solomon, H. M., and Lutz, M. F. (1987). “Mancozeb: Oral (Gavage) Developmental Toxicity Study in Rabbits.” Unpublished Rep. 86R021 from Rohm and Haas Company. Solomon, H. M., Lutz, M. F., and Kulwich, B. A. (1988). “Mancozeb: 2Generation Reproduction Study in Rats.” Unpublished Rep. 87R-020 from Rohm and Haas Company. Sortwell, R., Heywood, R., Allen, D., Prentice, D., and Cherry, C. (1977). “Metiram Preliminary Oral Toxicity Study in Rhesus Monkeys. Repeated Dosage for 4 Weeks.” Rep. BASF 77/0149, BSF 265/77264 from Huntingdon Research Centre, Huntingdon, UK. Unpublished report of BASF Corporation. Sortwell, R., Allen, D., Heywood, R., and Street, A. (1979). “Metiram (Containing 2.2% Ethylenethiourea) Oral Toxicity Study in Rhesus Monkeys (Repeated Dosage for 26 Weeks with Recovery Period).” Rep. BASF 79/0082, BSF 267/78263 from Huntingdon Research Centre, Huntingdon, UK. Unpublished report of BASF Corporation. Stadler, J. (1990). “Combined Chronic Toxicity/Oncogenicity 2-Year Feeding Study with Mancozeb in Rats.” Unpublished Rep. 259-89 of DuPont Haskell Laboratory. Stadler, J. (1991). “Neuropathology Study in Rats with Mancozeb.” Unpublished Rep. 217-89 of DuPont Haskell Laboratory. Steinmetz, K. A. and Potter, J. D. (1991). Vegetables, fruit, and cancer. I. Epidemiology. Cancer Causes Control 2, 325–357. Takayama, S., Aihara, K., Onodera, T., and Akimoto, T. (1986). Antithyroid effects of propylthiouracil and sulfamonomefhoxine in rats and monkeys. Toxicol. Appl. Pharmacol. 82, 191–199. Teramoto, S., Shingu, A., Kaneda, M., and Saito, R. (1978). Teratogenicity studies with ethylenethiourea in rats, mice and hamsters. Congenital Anomalies 18, 11–17. Terrill, J. B. (1990). “Acute Inhalation Toxicity Study with Maneb in the Rat.” Rep. HLA 2567-100 from Hazleton Laboratory, Vienna, VA. Unpublished report of Elf Atochem North America, Inc. Terrill, J. B. (1991). “A Four-Week Study in the Rat to Determine the Lung Tissue Residue and Effect upon the Thyroid Function with Maneb.” Rep. HLA 2567-101 from Hazleton Laboratory, Vienna, VA. Unpublished report of Elf Atochem North America, Inc. Tesh, J. M., McAnulty, P. A., Willoughby, C. R., Enticott, J., Wilby, O. K., and Tesh, S. A. (1988). “Mancozeb: Teratology Study in the Rat.” Rep. 87/PTC 007/365 from Life Science Research Ltd., Suffolk, UK. Unpublished report of Elf Atochem North America, Inc. Teshima, R., Nagamatsu, K., Kido, Y., and Terao, T. (1981). Absorption, distribution, excretion and metabolism of ethylenethiourea in guinea pigs. Eisei Kagaku 27, 85–90.

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Thomas, G. A. and Williams, E. D. (1991). Evidence for and possible mechanisms of non-genotoxic carcinogenesis in the rodent thyroid. Mutation Res. 248, 357–370. Tomlinson, H. L., and Longacre, S. L. (1988). “Mancozeb Dermal Absorption Study in Male Rats.” Unpublished Rep. 88R-218 of Rohm and Haas Company. Tompkins, C. E. (1992). “An 18-Month Dietary Oncogenicity Study in Mice with Maneb Technical.” Rep. WIL-134008 from WIL Research Laboratories, Inc., Ashland, OH. Unpublished report of Elf Atochem North America, Inc. Trutter, J. A. (1988a). “Subchronic Toxicity Study with Maneb Technical in the Rat.” Rep. HLA 153-140 from Hazleton Laboratories, Vienna, VA. Unpublished report of Elf Atochem North America, Inc. Trutter, J. A. (1988b). “A 21-Day Dermal Toxicity Study in Rabbits with Maneb Technical.” Rep. HLA 153-139 from Hazleton Laboratories, Vienna, VA. Unpublished report of Elf Atochem North America, Inc. Trutter, J. A. (1988c). “Mancozeb: 4-Week Repeat Dermal Toxicity Study in Rats.” Rep. HLA 417-432 from Hazleton Laboratories, Vienna, VA. Unpublished report of Rohm and Haas Company. Ulland, B. M., Weisburger, J. H., Weisburger, E. K., Rice, J. M., and Cypher, R. (1972). Thyroid cancer in rats from ethylenethiourea intake. J. Nat. Cancer Inst. 49, 583–584. Ullman, L., Sacher, R., Porricello, T., Luetkemeier, H., Vogel, W, Vogel, O., Wilson, J., and Terrier, H. (1987). “Report on the Subacute 21-Day Repeated Dose Dermal Toxicity Study with Polyram DF in Rabbits. Rep. BASF 87/0260, ZNT 86/314 from Research and Consulting Company AG, Ltd., Itingen, Switzerland. Unpublished report of BASF Corporation.

Hayes’ Handbook of Pesticide Toxicology

Ulrich, C. E. (1986a). “Thirteen-Week Subchronic Inhalation Toxicity Study on Maneb in Rats (Final Report).” Rep. 550-001 from International Research and Development Corporation, Mattawan, MI. Unpublished report of Elf Atochem North America, Inc. Ulrich, C. E. (1986b). “Report on the Thirteen Week Subchronic Inhalation Toxicity Study on Metiram in Rats.” Rep. RZ 86/407 and Addendum BASF 87/0414 from International Research and Development Corporation, Mattawan, MI. Unpublished report of BASF Company. Ulrich, C. E. (1987). “Thirteen-Week Subchronic Inhalation Toxicity Study on Maneb in Rats (Addendum to Final Report Covering Recovery Phase).” Rep. 550-001 from International Research and Development Corporation, Mattawan, MI. Unpublished report of Elf Atochem North America, Inc. Watts, M. H., and Chan, P. K. (1984a). “Dithane M-45: Acute Oral Toxicity Study in Rats.” Unpublished Rep. 83R-218 from Rohm and Haas Company. Watts, M. H., and Chan, P. K. (1984b). “Dithane M-45: Acute Oral Toxicity Study in Rats and Mice.” Unpublished Rep. 83R-213A and B from Rohm and Haas Company. Whysner, J., Ross, P. M., and Williams, G. M. (1996). Phenobarbital mechanistic data and risk assessment: enzyme induction, enhanced cellular proliferation, and tumor promotion. J. Pharmacol. Therapeutics 71, 153–191. World Health Organization (1994). “The WHO Recommended Classification of Pesticides by Hazard and Guidelines to Classification, 1994–1995.” International Program on Chemical Safety, WHO/PCS/94.2.

Chapter 79

Symmetrical Triazine Herbicides: A Review of Regulatory Toxicity Endpoints Charles B. Breckenridge1, J. Charles Eldridge2, James T. Stevens2 and James W. Simpkins3 1

Syngenta Crop Protection, Inc, Greensboro, NC, USA Wake Forest University School of Medicine, Winston-Salem, NC, USA 3 University of North Texas Health Science Center, Fort Worth, TX, USA 2

79.1  Introduction

79.1.1  Chemistry

Symmetrical triazines (s-triazine) have been used as selective herbicides in agriculture in the United States and other parts of the world for more than 50 years (Breckenridge et al., 2008). Certain of these s-triazine herbicides remain agronomically and commercially important in the food production system for the pre-emergent and post-emergent control of broadleaf weeds and certain grasses. The safety database of this class of chemistry has been reassessed by regulatory authorities around the world including the World Health Organization (JMPR, 2007), while ongoing reviews continue in some countries as regulations are updated and new studies become available. In this review chapter, the lowest observed adverse effect levels (LOAELs) and no observed adverse effect levels (NOAELs) were based upon published regulatory decisions such as US Environmental Protection Agency (USEPA) Reregistration Eligibility Decisions (REDS) and European Union reviews when available. When such assessments were not available, individual studies were consulted, if available to the authors, and the LOAELs and NOAELs reported by the study director were utilized. Throughout this document, doses are specified on a milligram/killigram body weight/day basis and standard conversion factors have been used to convert feeding levels typically expressed as part per million (ppm) concentrations in feed to mg/kg/day dose units. Additionally, as atrazine is the major product in this class of s-triazines in the United States, there is included extensive consideration of its regulatory position in regard to mammary tumorigenesis in the Sprague-Dawley rat the mode of action for these tumors development in rat, its non-relevance to humans, and the implications to risk assessment for atrazine.

The major commercial symmetrical triazines are further divided into chloro-s-triazines: simazine, atrazine, propazine, terbuthylazine, cyanazine; the thiomethyl-s-triazines: ametryn, prometryn, terbutryn; and the methoxy-s-triazine prometon. Symmetrical triazines (s-triazines) have a chlorine, sulfur, or oxygen atom at the 2-position of the ring and are usually substituted in the 4- and 6-positions with alkylamino-groups. Cyanazine (discontinued) contains a 2-cyano-isopropylamino-substituent at the 4-position on the ring. The structures of these symmetrical triazines are shown in Figure 79.1.

Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

79.1.2  Uses S-Triazines inhibit photosynthesis in organisms with oxygenevolving photosystems (Trebst, 2008). As shown in Fig. 79.2, s-triazines block photosynthetic electron transport by displacing plastoquinone from a specific-binding site on D1 protein subunit of photosystem II (PSII). This action is fully reversible when the exposure level is below the threshold of photosynthetic inhibition.

79.2  Hazard identification Hazard endpoints were identified for the s-triazine herbicides based upon protocols established under the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA) (U.S. EPA, 1982); hazard studies were conducted according to Good Laboratory Practices (U.S. EPA, 1979).

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1712

s-Chloro-triazine

s-Thiomethyl-triazine

Cl

N

S— N

N

N

N N Atrazine

N

N

N

S—

O—

N N Propazine

N

N N Prometryn

N N N Prometon

N

---

N

N

N N N Terbuthylazine Cl

N

N

S— N

N

N

N

Cl

N

---

N N Simazine Cl

N

--N

N

N

---

N

N N Ametryn

Cl

N

Figure 79.1  Structures of the Photosynthesis Inhibiting Triazine Herbicides.

s-methoxy-triazine

N N Terbutryn ---

---

Photosystem II

Cytochrome b,f complex

N

N N Cyanazine

C≡N

D2

PQA

PQB

D1

s-triazines bind to the D1 protein complex of PS II interupting electron transfer

cyt b

FSA FSB

PQ

FSX

cyt b

Phe0 P680

Vitamin K1

PQH2 Y2 D2

Mn Mn Mn Mn 33 23 16 O2+4H 2H2O

H+ ATP

Ferrodoxin

PQH2

D1

ATP synthase ADP+P1

Stroma

2H+

Antenna complex

Photosystem I

FNR NADP+ Internal membrane

chorophyI a

FeS cyt f

P700 PC (2H+)

Lumen Represents electron transfer Represents transfer of protons

D1 = protein D1, D2 = protein D2 33. 16 and 23 are other proteins of reaction centre II

Figure 79.2  Linear electron transport system of photosynthesis. Photosystem (from Swindell, 2003).

The Food Quality Protection Act (FQPA) of 1996 reauthorized FIFRA provisions and requires tolerances to be reassessed as part of a re-registration process (Bliley, 1996). FQPA provides for a single, health-based standard, and eliminated

the problem posed by having different standards for pesticide concentration in raw and processed foods. FQPA required that in the process of setting tolerances for pesticide residues in food, the USEPA must evaluate the aggregate risk arising

Chapter | 79  Symmetrical Triazine Herbicides

from exposure to pesticides from all routes of exposure including oral, dermal or inhalation exposure.

79.2.1  Acute Studies Acute toxicity generally applies to effects that result from a single dose or single exposure of a chemical following oral, dermal, or by inhalation exposure. The values calculated through the oral and dermal routes of exposure are referred to as Lethal Dose 50 (LD50), the amount of chemical required to kill 50% of the test animals in a group within the first 14 days following exposure. Lethal Concentration 50 (LC50) mortality is equivalent to the concentration of chemical administered (usually as an aerosol) leading to 50% mortality of the animals by inhalation exposure. Acute studies are also conducted to evaluate the irritation potential of the chemical after application to skin and eyes and are used to establish the personal protective equipment, re-entry interval, and signal words for product labels for all crop protection chemicals (Sumner et al., 1995). The commercially important triazine herbicides are relatively nontoxic, and they are not generally irritating to the skin or eye (Table 79.1).

79.2.2  Toxicity after Repeat Exposure The United States Environmental Protection Agency guidelines (U.S. EPA, 1982, 1998a) for subchronic studies for

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pesticides specify treatment of rats, mice, or dogs with the chemical for specified lengths of time. Rat and mouse studies are 90 days in duration, and lifetime studies are typically 24 and 18 months, respectively. In dogs, the studies are usually conducted for 90 days, 1 year, or 2 years. In all cases, animals are divided into test groups of typically 10 to 50 rats or mice and 4 to 6 dogs. At least four test groups are used in each study, with one group receiving no chemical (controls) and three groups receiving low, medium, or high concentrations of the chemical in their diets. Urinalysis, hematology, and clinical chemistry parameters are evaluated and gross and microscopic pathological examinations are performed on up to 50 tissues. Maximally tolerated doses (MTD) are tested in order to demonstrate toxicity (up to 1000 mg/kg/ day in the diets). The MTD is defined as the highest concentration of test chemical that can be administered without severe toxicity; often a 10% reduction in body weight gain has been used as a default criterion for establishing the MTD (Farber, 1987; Foran et al., 1997). With this approach, it is possible to determine whether a chemical damages or alters any organ or tissue at high doses, and to establish levels of the chemical which produce no observable adverse effects (NOAEL), and the lowest level at which effects are noted (the LOAEL). The responses of repeated daily exposure of rats and dogs to the selected triazine herbicides are presented in Table 79.2. The resulting doses and endpoints are expressed in milligrams of chemical per kilogram of body weight per day of dosing.

Table 79.1  EPA Acute Hazard Classification of the Technical Grade for Selected Triazine Herbicidesa, b Triazine technicala

Eye irritation

Skin irritation

Oral LD50 (mg/kg)

Dermal LD50 (mg/kg)

Inhalation Sensitization LC50 potential (mg/L)

Signal word

Group

Chemical

s-Cl

Atrazine

Nonirritating

Nonirritating

1869–3090

2000

5.8

Nonsensitizerc

Caution

Simazine

Slight irritation

Mild irritation

5000

2000

1.7

Nonsensitizer

Caution

Propazine

Slight irritation

Nonirritating

5050

5050

1.22

Nonsensitizer

Caution

Terbuthylazine

Moderate irritation

Mild irritation

1000–1590

2000

5.3

Nonsensitizer

Caution

Cyanazine

Nonirritating

Nonirritating

182–334

2000

0.81

Nonsensitizer

Warning

Ametryn

Mild irritation

Nonirritating

1009–1590

2020

5.2

Sensitizer

Caution

Prometryn

Slight irritation

Mild irritation

1802–2076

3170

4.96

Nonsensitizer

Caution

Terbutrynd

Nonirritating

Nonirritating

2450

2000

2.2

Nonsensitizer

Caution

Prometon

Slight Irritation

Nonirritating

1520–4350

2020

3.2

Nonsensitizer

Caution

s-SCH3

s-OCH3 a

This table lists only technical products; formulated products used for agriculture may have more restrictive labelling due to the formulants used. Values for acute hazard from EPA Registration Eligibility Documents for each product or/and Breckenridge et al., 2008. c Sensitizer in the Guinea Pig maximum test but not the human Buehler test. d Studies were conducted using an 80% formulation of the active ingredient (80 W). b

Hayes’ Handbook of Pesticide Toxicology

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With the exception of rats and dogs fed terbuthylazine or cyanazine, the NOAEL values were generally 2.5 mg/kg/day or higher, and LOAEL values were 14 mg/kg/day or higher. The most common observation was not a change in a specific organ or tissue, but a reduction in body weight gain.

79.2.3  Developmental and   Reproductive Toxicity Hazard testing also includes the examination of the potential of a chemical to affect the development of offspring and the determination of whether it induces birth defects in either rats or rabbits. These tests have historically been called

teratology studies, but are now usually referred to as developmental toxicity studies. In addition to developmental toxicity studies, a reproduction study in rats that involves feeding diets containing the chemical to young adult male and female rats for approximately 3 months prior to mating is conducted. The females are allowed to produce a litter of offspring that are then reared to adulthood. The animals are fed diets containing the chemical during this entire period of time. After reaching sexual maturity, the second-generation animals are allowed to mate. The results of such studies conducted with the selected s-triazine herbicides are presented in Table 79.3. None of the s-triazine herbicides, with the exception of cyanazine, produced developmental or reproductive effects. Cyanazine caused developmental effects in the rat and

Table 79.2  Hazard Assessment for Repeat Exposure to the Selected Triazine Herbicides Triazine

Species/study

Group

Chemical

s-Cl

Atrazine c

Simazine

Propazine

Terbuthylazine

s-SCH3

Ametryn

Prometryn

f

Terbutryn

s-OCH3

a

Prometon

LOAELb

Rat/90-day oral

3.3

34.5

↓ Body wt. Gain

Dog/52-week oral

4.97

33.7

Heart/myocardium

Rat/90-day oral

14.3

14.3

↓ Body wt. Gain

Dog/90-day oral

6.9

64

↓ Body wt. Gain

Rat/90-day oral

13

50

↓ Body wt. Gain

Dog/90-day oral

7

25

↓ Body wt. Gain

Rat/28-day oral

2.3

2.3

↓ Body wt. gain, ↑organ weight

d

2.1

7.1

↓ Body wt. gain, chances in hematological and clinical chemistry

Rat/90-day oral (Study 2) e

4.0

8.0

↓ Body wt. Gain

Dog/52-week oral

0.4-1.25

1.25–7.8

↓ Body wt. Gain

Rat/90-day oral

2.5

5.0

↓ Body wt. Gain

Dog/52-week oral

0.7

3.0

Hematological effects

Rat/90-day oral

7.4

36

Hematological effects

Dog/52-week oral

7.2

70

Liver

Rat/90-day oral

2.5

25

↓ Body wt. Gain

Dog/104-week oral

3.7

37.5

Liver, kidney, bone marrow

Rat/90-day oral

50

140

↓ Body wt. Gain

Dog/26-week oral

10

25

Stomach

Rat/90-day oral

5

15

↓ Body wt. Gain

Dog/52-week oral

5

20

↓ Body wt. Gain

No observable adverse effect level. Lowest observable adverse effect level. c USEPA utilized the atrazine chronic NOEL of 1.8 mg/kg/day for cumulative risk assessment. d UK 2007 draft review submitted to the European Commission. e USEPA 1995 Registration Eligibility Decision. f Studies were conducted using an 80% formulation of the active ingredient (80 W). b

Effects used to establish the LOAEL

NOAELa

Rat/90-day oral (Study 1)

Cyanazine

mg/kg/day

Chapter | 79  Symmetrical Triazine Herbicides

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Table 79.3  Results of Rat and Rabbit Developmental and a Two-Generation Rat Reproduction Study with s-Triazine Herbicides Triazine Group

Chemical

s-Cl

Atrazine

Simazine

Propazine

Terbuthylazine

Cyanazine

s-S-CH3

Ametryn

Prometryn

Terbutryn i

s-OCH3

a

Prometon

Study/species

Developmental/ reproduction

Toxicity observed

Developmental/rat

None

Developmental/rabbit

HDTa

LOAELb

NOAELc

↓ Body wt. gain

700

70

10

None

↓ Body wt. gain

75

75

5

Reproductive/rat

None

↓ Body wt. gain

35

35

3.8

Developmental/rat

None

↓ Body wt. gain

600

300

30

Developmental/rabbit

None

↓ Body wt. gain

200

75

5

Reproductive/rat

None

↓ Body wt. gain

29–35

6

0.6

Developmental/rat

None

↓ Body wt. gain

600

100

10

Developmental/rabbit

None

↓ Body wt. gain

50

10

2

Reproductive/rat

None

↓ Body wt. gain

50

50

5

Developmental/rat d

None

↓ Body wt. gain

30

30

5

Developmental/rabbit d

None

↓ Body wt. gain

4.5

4.5

4.5

Reproductive/rat (Study 1) e

Reduced fertility

↓ Body wt. gain

20–26

4.5–26

0.4 f–4.5 g

Reproductive/rat (Study 2) e

None

↓ Body wt. gain

25–36

7.3–10.4

3.6–4.5h

Reproductive/rat (Study 3) e

None

↓ Body wt. gain

14.6–18.1 7.1–11.4

3.5–4.5h

Developmental/rat

Positive

↑ Malformations

75

5

5

Developmental/rabbit

Positive

↑ Malformations

4

2

1

Reproductive/rat

None

↓ Body wt. gain

15

5

1.5

Developmental/rat

None

↓ Body wt. gain

250

50

5

Developmental/rabbit

None

↓ Body wt. gain

60

60

10

Reproductive/rat

None

↓ Body wt. gain

131

131

13

Developmental/rat

None

↓ Body wt. gain

250

250

50

Developmental/rabbit

None

↓ Body wt. gain

72

72

12

Reproductive/rat

None

↓ Body wt. gain

96.7

47.8

0.6

Developmental/rat

None

↓ Body wt. gain

500

500

50

Developmental/rabbit

None

↑ Unossified sternabra

75

75

10

Reproductive/rat

None

↓ Body wt. gain

150

150

15

Developmental/rat

None

↓ Body wt. gain

360

120

36

Developmental/rabbit

None

↓ Body wt. gain

24.5

24.5

3.5

Reproductive/rat

None

↓ Body wt. gain

75

25

1

Highest dose tested. Lowest observable adverse effect level. No observable adverse effect level. d USEPA 1995 Reregistration Eligibility Decision Document. e UK 2007 draft review submitted to the European Commission. f Maternal NOAEL. g Reproductive NOAEL. h Defined by the European Commission as based on parental toxicity. i Studies were conducted using an 80% formulation of the active ingredient (80 W). b c

mg/kg/day

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1716

rabbit at the highest doses tested. Effects noted at doses that were toxic to the mothers were cyclopia and diaphragmatic hernia in rabbits, and an apparent increase in the incidence of skeletal variations (i.e., anomalies) in rats (U.S. EPA, 1994).

79.2.4  Mutagenicity Weisburger (1975) noted that certain chemical carcinogens are capable of interacting directly with genetic material such as DNA. Thus, several genotoxicity studies aimed at identifying the alteration of genetic material or mutation were introduced into hazard testing. These include tests to examine the possible interaction with genes (gene mutation tests), interaction with the chromosome (clastogenic tests), and direct interaction with DNA (classified as other tests). The results for the selected triazine herbicides are presented in Table 79.4. All of the triazines were negative in the studies in Table 79.6. The overall mutagenic potentials of atrazine, simazine, and cyanazine have been reviewed by Brusick (1994), Hauswirth and Wetzel (1998), Bogdanffy et al. (2000), and the USEPA (2003a), respectively. The weight of the evidence indicates that these s-chloro-triazines are not mutagenic a conclusion reached by the USEPA (2003a, 2006a) and the JMPR (2007). Cyanazine showed only limited evidence of mutagenicity (USEPA, 1994).

79.2.5  Oncogenicity Assessment Since individuals may be exposed to low levels of chemicals over a portion of their lifespan, studies to evaluate

lifetime exposures are conducted in animal bioassays. An important aspect of these studies is the evaluation of the potential of a compound to increase the incidence of benign or malignant tumors and/or to accelerate time of onset of tumors). For the laboratory studies, mice and rats are divided into at least three treatment groups and a control group, with a minimum of 50 animals of each sex in each group. The groups of animals are fed selected concentrations of the test chemical in their diet for 18 months and 24 months, respectively. The levels of the test chemical administered in the diet are generally chosen from repeat dose feeding studies that-are at least 90 days in duration and are normally used to establish the NOAEL, LOAEL, and MTD. Following lifetime feeding studies at the prescribed treatment levels, veterinary-pathologists examine approximately 50 tissues from each animal for the presence of tumors or other evidence of tissue damage. The results of such oncogenicity studies in mice, conducted with the s-triazine herbicides, are presented in Table 79.5. None of these s-triazines showed any evidence of inducing tumors in mice. The highest doses ranged from 87 to 1140 mg/kg/day and were equal to or exceeded the MTD. In rats, the chloro-s-triazines (e.g., atrazine, cyanazine, propazine, and simazine) resulted in either an increased incidence or an earlier onset of mammary tumors when administered to female Sprague-Dawley (SD) rats at high feeding levels, as presented in Table 79.6. The thiomethyl- and methoxy-s-triazines were not carcinogenic in the SD rats at all feeding levels including those exceeding the MTD; the exception was terbutryn,

Table 79.4  Mutagenicity Studies with the Selected s-Triazine Herbicides Triazine

Gene mutation

Clastogenic micronucleus

Other

Group

Chemical

Ames

E. Coli REC

Mouse lymphoma

s-Cl

Atrazine

Negative

Negative

Negative

Negative

Negative

Negative

Simazine

Negative

—

Negative

Negative

Negative

—

Propazine

Negative

Negative

Negative

Negative

Negative

—

Terbuthylazine

Negative a

Negative

Negative

Negative b

Negative

—

Terbuthylazine

Negative

__

Negative

Negative

Negative

__

Cyanazine

Negative

—

Negative

Negative

Negative

—

Ametryn

Negative

—

Negative

Negative

Negative

Negative

Prometryn

Negative

Negative

Negative

Negative

Negative

—

Terbutryn

Negative

—

—

Negative

Negative

Negative

Prometon

Negative

—

—

Negative

Negative

—

s-SCH3

s-OCH3 a

Also negative in Chinese hamster ovary/HGPTR assay. Also negative in the Chinese hamster ovary.

b

DNA repair Dominant lethal

Chapter | 79  Symmetrical Triazine Herbicides

1717

where an increased incidence of mammary, thyroid and liver tumors were observed in female SD rats at feeding levels that exceeded the MTD.

79.3  Mode of action for mammary tumor formation in the spraguedawley rat at high doses The Sprague-Dawley is one of the most commonly used laboratory rat strains. However, it has limitations when used to evaluate the effects of chemicals on the endocrinerelated tumors of a high spontaneous tumor incidence in the pituitary and mammary gland. At approximately 9–12 months of age, the female SD rat begins to experience prolonged periods of vaginal estrus (Eldridge et al., 1996; Eldridge et al., 1999; Sirnpkins et al., 1998). Most laboratory rats (including SD rats prior to approximately 9 months) spend about 20-25% of their time in estrus. The SD rat spends increasing amounts of time in estrus after this period, often 40% by 12 months of age, and in some

cases achieves persistent estrus at senescence (Eldridge et al., 1998). This unique physiology derives from deficient neuroendocrine control of the secretion of gonadotropinreleasing hormone (GnRH) from the hypothalamus. With decreasing release of GnRH, pituitary surges of luteinizing hormone (LH) gradually decreases until it is inadequate to stimulate ovulation. Thus, the female SD rat cease to go through their normal estrous cycle and become ‘stuck in estrus.’ This is represented in Figure 79.3. These prolonged periods of estrus caused the SD rats to experience prolonged exposure to estrogen and prolactin produced by the ovary and pituitary, respectively (Simpkins et al., 1998). Both prolactin and estrogen are known to increase the incidence of mammary tumors in rats (Cutts and Noble. 1964). Ovulatory failure in the SD rat results in the mammary gland becoming hyper-stimulated by estrogen arising from the ovarian follicle and prolactin produced by the pituitary; this overstimulation of the pituitary also results in a high incidence of pituitary adenomas (Stevens et al., 1994). Over a sufficiently long time period, this hyperstimulation translates into a proliferative response in the

Table 79.5  Carcinogenicity Studies in the Mouse Conducted with Selected s-Triazines Triazine Group

Herbicidea

s-Cl

Atrazine

s-SCH3

c

Reference

LOAELd

Negative

386e

43

194–247

↓ Body weight gain and thrombi in both sexes

Hauswirth and Wetzel (1998)

Simazine

Negative

542–652e

5.3

132

↓ Body weight gain in both sexes

Hauswirth and Wetzel (1998)

Propazine

Negative

450e

15

450

Cardiac fibrosis and focal degeneration

IRIS (1997a)

Terbuthylazine

Negative f

87–89 e

17 g

87–89

↓ Body weight gain in both sexes; hematologic changes in males

USEPA (1995)

Terbuthylazine

Negative \h

99–118

14.6–15.5 g 37–40

↓ Body weight gain

UK (2007)

Cyanazine

Negative

143

1.4

3.6

↓ Body weight gain in both sexes

USEPA (1994)

300

300

↓ Body weight gain in both sexes

USEPA (2004)

143

429

↓ Body weight gain in both sexes

USEPA (1996)

429

429

No effects observed

Jessup (1980)

60

570

↓ Survival, kidney necrosis, hepatocellular hypertrophy, splenic atrophy

Osheroff (1988)

Ametryn

Prometon

i

Negative

300

Negative

429

Negative Negative

e

e

429

1140

e

CD1 mice were tested for all chemical except terbuthylazine where Tif: MAGf was used. Highest dose tested. No observable adverse effect level d Lowest observable adverse effect level. e Maximum tolerated dose exceeded. f USEPA 1995 Reregistration Eligibility Decision Document g Draft European Commission review defined this dose as a NOAEL h UK 2007 draft review submitted to the European Commission. i Studies were conducted using an 80% formulation of the active ingredient (80 W). b

Other effects

NOAELc

Terbutryn

a

Feeding level (mg/kg/day) HDTb

Prometryn

s-OCH3

Cancer potential

Hayes’ Handbook of Pesticide Toxicology

1718

Table 79.6  Carcinogenicity Studies in the Rat with Selected s-Triazines Triazine

Other effects

Reference

NOAELb LOAELc

Mammary (SD) d

50 e

2.5

20

↓ Body weight gain in both sexes

USEPA (2003a,b)

Negative (F-344)d

20

4.8

20

↓ Body weight gain in both sexes

USEPA (2003a,b)

Simazine

Mammary (SD)

45.8–63.1e

0.5

4.2–5.3

↓ Body weight gain in both sexes; hematologic effects and ↑ mortality in females

USEPA (2006b)

Propazine

Mammary (SD)

50e

5.8

50

↓ Body weight gain in both sexes

Stevens et al. (1994)

Terbuthylazine

Mammary, Leydigf Study 1 (Tif:RAI)d

42–53e

1.2

1.2

↓ Body weight gain in both sexes

USEPA (1995)

Negativeg, Study 2 (Wistar) d

1.6

0.35

1.6

↓ Body weight gain

UK (2007)

Mammaryg, Study 3 (Wistar)

5.5–7.6

0.4

1.7

↓ Body weight gain

UK (2007)

Cyanazine

Mammary (SD)

2.5e

0.2

1.0

↓ Body weight gain in females; hyperactivity in the males

U.S. EPA (1994)

Ametryn

Negative (SD)

145–176e

21

145

↓ Body weight gain in both sexes; hematological effects in females

USEPA (2004)

Prometryn

Negative (SD)

75e

29

61

↓ Body weight gain in both sexes

USEPA (1996)

Terbutrynh

Mammary, Liver, Thyroid (SD)

150e

0.1

15

↓ Body weight gain in both sexes

Stevens et al. (1994)

Prometon

Negative (SD)

75

1

25

↓ Body weight gain in both sexes

Stevens et al. (1994)

Herbicide

s-Cl

Atrazine

s-OCH3

Feeding level (mg/kg/day) HDT a

Group

s-SCH3

Tumor response

a

Highest dose tested. No observable adverse effect level. c Lowest observable adverse effect level. d Strain of rat tested (SD  Sprague-Dawley, Tif:RAI  Sprague-Dawley derived strain of rat, F-344  Fischer 344, Wistar  Wistar) e Maximum tolerated dose exceeded. f USEPA 1995 Reregistration Eligibility Decision Document (elevated incidence at doses that exceed the MTD; Inadequate evidence. Classified as Category D). g UK 2007 draft review submitted to the European Commission. h Studies were conducted on an 80% formulation of active ingredient (80 W). b

Figure 79.3  Schematic of Normal and Constant Estrus in Sprague-Dawley Rats.

Chapter | 79  Symmetrical Triazine Herbicides

mammary gland characterized by the development of a high spontaneous incidence of adenocarinoma (high estrogen, moderate prolactin levels) or fibroadenoma (high prolactin with a background of estrogen) (Stevens et al., 1994; Stevens et al., 1999; Sielken et al., 2005). The reproductive aging process normally occurring in the female SD rat is strain specific. The Fischer 344 rat, does not display this neuroendocrine deficiency and do not have a high spontaneous incidence of mammary or pituitary tumors (Eldridge et al., 1998). Detailed studies have shown that Fischer 344 rats administered high doses of atrazine do not develop either an increased incidence or an early onset of mammary tumors (Thakur et al., 1998; Wetzel et al., 1994), unlike the findings noted in similarly treated female SD rats (Stevens et al., 1994; Wetzel et al., 1994; Hauswirth and Wetzel, 1998). Furthermore, when the major internal source of estrogen was removed from the female SD rats by surgical removal of the ovaries at the beginning of the study, no mammary tumors were found after 2 years of atrazine treatment (Stevens et al., 1999). Examination of the reproductive cycles of intact female SD rats fed high doses of atrazine over their lifetimes showed that prolonged periods of estrus occurred earlier in the treated group than in the control group (Hauswirth and Wetzel, 1998). Subsequent studies showed that high doses of atrazine administered to female SD rats reduced the magnitude of the LH surge, resulting in a failure of ovulation to occur (Simpkins et al., 1998). Low doses of atrazine had no effect on the LH surge, the estrous cycle, or the time to appearance or incidence of mammary tumors (Simpkins et al., 1998). These results indicate that even in female SD rats there is a threshold dose below which there are no adverse effects on reproductive processes. Finally, when high-dose atrazine-treated animals were given GnRH, the hormone that is responsible for triggering the LH surge, the LH surge was restored. This finding suggests that the pituitary LH releasing mechanisms function normally in atrazine-treated animals (Cooper et al., 1995). In summary, high doses of chloro-s-triazines appear to accelerate the development of mammary tumors in the SD rat; this phenomenon occurs in a strain of rat that is already prone to spontaneously developing mammary tumors because of an inherent age-dependent deficiency in the regulation of the estrous cycle. The earlier appearance of mammary tumors in female SD rats treated with high doses of atrazine is attributed to an increased exposure to endogenous estrogen and prolactin, secondary to the lengthening of the estrous cycle. Removal of endogenous estrogen in female SD rats by ovariectomy prevents the appearance of mammary tumors, even in animals that have received high doses of atrazine for 2 years. In the SD female there is a dose of atrazine (approximately 2.5 mg/kg/day) that has no effect on the estrous cycle or

1719

mammary tumor incidence and/or onset. The mammary tumor response to high doses of atrazine is unique to the female SD rat; the response was not observed in male or female CD-1 mice, Fischer 344 rats, or male SD rats. As noted below, the mode of action determined in the SD female rat is not relevant to humans (IARC, 1999; USEPA, 2000, 2002, 2003a, b; JMPR, 2007).

79.4  Epidemiology Several reviews of the epidemiological evidence relating to atrazine exposure and the occurrence of cancer in humans have been conducted by Loosli (1995), Neuberger (1996), and Sathiakumar and Delzell (1997). The weight of the evidence indicates there is no basis for concluding that there is a causal association between exposure to atrazine and cancer in humans. Cohort studies conducted at a production facility over a long period of follow-up have not identified any increased cancer risk, including the risk of non-Hodgkin’s lymphoma (MacLennan et al., 2002). The number of prostate cancer cases at a production facility was elevated but was fully accounted for by the prostate cancer screening bias operative at the plant as a result of an advanced medical surveillance program (Hessel et al., 2004). The null results reported for prostate cancer in the large cohort of licensed pesticide applicators who are members of the US governmentsponsored Agricultural Health Study further support this conclusion (Alavanja et al., 2003; Rusiecki et al., 2004). In fact, Blair et al. (2005) concluded that, ‘No exposure and response gradient was noted for any cancer among farmers exposed to atrazine, including prostate.’ A review of the case-control studies principally on nonHodgkin’s lymphomas, has not established a causal association between atrazine use and the occurrence of this disease (De Roos et al., 2003). This conclusion has also been reached in numerous authoritative reviews (USEPA, 2003a. b; International Agency for Research on Cancer (IARC), 1999; United Kingdom (UK), 2000; Australian Pesticides and Veterinary Medicines Authority (APVMA, 2004, 2008). The ecological epidemiology studies, which do not measure exposure or disease at the level of the individual, have generated null relationships, inverse relationships (Van Leeuwen et al., 1999) and a few positive associations that have not been supported by the results from cohort studies (Mills, 1998, 2003; Muir et al., 2004). The results from some of the studies have been contradictory (e.g., Kettles et al., 1997 versus Hopenhayn et al., 2002) or considered implausible (Van Leeuwen et al., 1999). A more definitive cohort study in Iowa and North Carolina failed to show any association between atrazine exposure and breast cancer among the wives of Agricultural Health Study workers (Engel et al., 2005).

1720

79.5  Chlorotriazine cancer classification A review of the mode of action data underlying the mammary tumor response observed in female SD rats treated with high doses of chlorotriazines has been reported elsewhere (Wetzel et al., 1994; Eldridge et al., 1996; Eldridge et al., 1999; Stevens et al., 1999; Eldridge and Wetzel, 2008). Upon review of the available atrazine oncogenicity database, the World Health Organization’s International Agency for Research on Cancer (IARC, 1999) concluded “there is strong evidence that the mechanism by which atrazine increases the incidence of mammary gland tumours in Sprague-Dawley rats is not relevant to humans.” Subsequently based upon a weight of the evidence analysis, USEPA and its Science Advisory Panel (USEPA, 2000) arrived at this same conclusion that this mode of action for mammary tumor formation in female SD rats is not relevant to humans. Alternate modes of action have also been considered and discounted by USEPA. The weight of the evidence indicates that atrazine and the chlorotriazines are not genotoxic (Brusick, 1994), estrogenic (Eldridge et al., 1999; Eldridge et al., 2007) or causes increased estrogen levels in vivo by upregulation of aromatase gene regulation (JMPR, 2007). Reviews from the USEPA (2003a), the European Union (UK, 2000), the APVMA (1997, 2004, 2008), IARC (1999), and JMPR (2008) concluded that the mechanism underlying the occurrence of atrazine-induced mammary tumors in female SD rats is not relevant to humans. Atrazine is classified as not likely to be a human carcinogen by the USEPA (2002). Based upon these results, it is concluded that:   1. The chloro-s-triazines accelerate the onset timing of mammary tumors in the female SD rat, a strain of rat that is already prone to developing mammary tumors spontaneously because of an inherent age-dependent deficiency in the regulation of the estrous cycle.   2. The earlier appearance of mammary tumors in female SD rats treated with high doses of atrazine is attributed to an increased exposure to endogenous estrogen and prolactin, secondary to the lengthening of the estrous cycle.   3. Removal of endogenous estrogen in female SD rats by ovariectomy prevents the appearance of mammary tumors, even in animals that have received high doses of atrazine.   4. In the SD female there is a lifetime dose of atrazine (2.5 mg/kg) that has no effect on the estrous cycle or mammary tumor incidence and/or onset.   5. The mammary tumor response to high doses of atrazine is unique to the female SD rat and is not observed in male SD rats, three strains of mice, or in the F-344 rat.   6. The effects of atrazine on the reproductive aging processes observed in female SD rats are not relevant

Hayes’ Handbook of Pesticide Toxicology

to humans. In women, reproductive senescence is characterized principally as an ovarian failure with a decrease in endogenous estrogen exposure at menopause, not the increase that is characteristic of the female SD rat in a state of persistent estrus.   7. The International Agency for Research on Cancer (IARC. 1999) concluded there is strong evidence that the mechanism by which atrazine increases the incidence of mammary gland tumors in SD rats is not relevant to humans and that atrazine is not classifiable as to its carcinogenicity to humans.   8. After review by its Scientific Advisory Panel (SAP), the USEPA (2000, 2002, 2003a) has concurred with IARC (1999) that the mammary tumor response observed in the SD female rat is not considered relevant to humans and USEPA classified atrazine as not likely to be a human carcinogen.   9. In addition, other regulatory bodies around the world have reviewed the data on atrazine and arrived at the same conclusion as IARC and USEPA. The APVMA (2004) concludes that published epidemiological data provides support for the absence of carcinogenicity potential for atrazine,’ and maintained an earlier (1997) conclusion that animal data on the carcinogenicity of atrazine has no relevance to humans (APVMA, 1997). The Australian Pesticides and Veterinary Medicines Authority (2008) again concluded that the published epidemiological data provided no support for any carcinogenicity potential of atrazine (APVMA, 2008). The UK’s Rapporteur Monograph on Atrazine (2000) conducted for the European Union concluded that the ‘classification of atrazine as a carcinogen is not appropriate. In 2001 the French Toxicity Research Commission on Pesticide Products cited the IARC, the USEPA, and the EU conclusion that there is an absence of carcinogenic effects of atrazine in humans (French Republic Ministry of Agriculture, 2001). 10. The USEPA’s Office of Pesticide Programs concluded again in the October 31, 2003 Interim Reregistration Eligibility Decision (IRED) that ‘considering the animal data and the human epidemiological data, atrazine is “not likely to be carcinogenic in humans”’ (USEPA, 2003b). 11. After a second USEPA Scientific Advisory Panel reviewed additional data on prostrate cancer in July 2003, USEPA’s revised IRED (October 31, 2003) concludes that ‘the Agency did not find convincing evidence of an association between triazines or atrazine and cancer’ (USEPA, 2003b). 12. The lack of relevance of these data to humans is supported by 50 years of manufacturing and use history for atrazine and other s-triazine herbicides. There is no epidemiological evidence linking atrazine exposure to any human health effects (Sathiakumar et al., 1992; Loosli, 1995; Neuberger, 1996).

Chapter | 79  Symmetrical Triazine Herbicides

13. In addition, publications in 2003, 2004, and 2005 from a recent large-scale government study show no association of atrazine and cancer (Alavanja et al., 2003; Blair et al., 2005; Engel et al., 2005; Rusiecki et al., 2004). 14. The Report of the Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment and the WHO Core Assessment Group on Pesticide Residues met in Geneva, Switzerland in 2007. In their report, “Pesticide Residues in Food 2007” they concluded that “The published epidemiological data provide no support for any carcinogenicity potential of atrazine” (JMPR, 2007).

79.6  Overall hazard assessment Evaluation of hazard profiles for the currently commercial products triazine herbicides reveals that these products are generally not acutely toxic, are well-tolerated when administered to animals over a long duration of time, are generally not developmental or reproductive toxins, and are not mutagenic or carcinogenic in mice or male rats. The chloro-s-triazines appear to produce an earlier onset or an increased evidence of mammary tumors in female SD rats at high doses. Because of the unique nature of reproductive aging in female SD rats, the carcinogenic response in this strain is not relevant for human risk assessment. Furthermore, a review of the epidemiological studies on atrazine does not support an association of exposure to atrazine and the occurrence of cancer.

References Alavanja, M. C. R., Samanic, C., Dosemeci, M., Lubin, J., Tarone, R., Lynch, C. F., Knott, C., Thomas, K., Hoppin, J. A., Barker, J., Coble, J., Sandler, D. P., and Blair, A. (2003). Use of agricultural pesticides and prostate cancer risk in the agricultural health study cohort. Am. J. Epidem. 157(9), 800–814. APVMA (formerly the National Registrations Authority for Agriculture and Veterinary Chemicals). (1997). The NRA Review of atrazine, November 1997. APVMA (2004). The reconsideration of approvals of the active constituent atrazine. Registrations of products containing atrazine and their associated labels. October 2004. APVMA (2008). Atrazine. Final Review Report and Regulatory Decision. Volume 1. March, 2008. http://www.apvma.gov.au/chemrev/downloads/atrazine_finalMay08.pdf Blair, A., Sandler, D., Thomas, K., Hoppin, J. A., Kamel, F., Coble, J., Lee, W. J., Ruslecki, J., Knott, C., Dosemeci, M., Lynch, C. F., Lubin, J., and Alavanja, M. (2005). Disease and injury among participants in the agricultural health study. J. Agri. Safety Health 1l(2), 141–150. Bliley, R. (1996). Food Quality Protection Act of 1996. 104 Congress, 2nd Session. Report 104-669, part 2, pp. 1–89. Government Printing Office, Washington, DC. Bogdanffy, M. S., O’Connor, J. C., Hansen, J. F., Gaddamidi, V., Van Pelt, C. S., and Cook, J. C. (2000). Chronic toxicity and

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oncogenicity bioassay in rats with the chloro-S-triazine herbicide cyanazine. J. Toxicol. Environ. Health Accepted for publication. Breckenridge, C. B., Werner, C., Stevens, J. T. and Sumner, (2008). Hazard Assessment for selected Symmetrical and Asymmetrical Triazine Herbicides. Chapter 26. In “The Triazine Herbicides” (H.M. Le Baron, McFarland, O. Burnside, and R. Clark, eds.), pp. 346–358. Brusick, D. J. (1994). An assessment of the genetic toxicity of atrazine: Relevance to health and effects. Mutation Res. 317, 133–144. Cooper, R. L., Parrish, M. B., McElroy, W. K., Rehnberg, G. L., Hein, J. F., Goldmann, J. M., Stoker, T., and Tyrey, L. (1995). Effect of atrazine on the hormonal control of the ovary. The Toxicologist 15, 294. Cutts, J. H. and Noble, R. L. (1964). Estrone-induced mammary tumors in the rat. II. Effect of alteration in hormonal environment on tumor induction, behavior, and growth. Cancer Res. 24, 1124–1130. De Roos, A. J., Zahm, S. H., Cantor, K. P., Weisenburger, D. D., Holmes, F. F., and Burmeister, L. L. (2003). Integrative assessment of multiple pesticides as risk factors for non-Hodgkin’s lymphoma among men. Occup. Environ. Med. 60, 1–9. Eldridge, J. C., Stevens, J. T., Wetzel, L. T., Tisdel, M. O., Breckenridge, C. B., McConnell, R. F., and Simpkins, J. W. (1996). Atrazine: Mechanisms of hormonal imbalance in female SD rats. Fund. Appl. Toxicol. 24(12), 2–5. Eldridge, J. C., McConnell, R. F., Wetzel, L. T., and Tisdel, M. O. (1998). Appearance of mammary tumors in atrazine-treated female rats: Probable mode of action involving strain-related control of ovulation and estrous cycling Chap. 32. In “Triazine Herbicides: Risk Assessment,” (L. G. Ballantine, J. E. McFarland, and D. S. Hackett, eds.), pp. 414–423. American Chemical Society, Washington, DC. Eldridge, J. C., Wetzel, L. T., and Tyrey, L. (1999). Estrous cycle patterns of Sprague–Dawley rats during acute and chronic atrazine administration. Reprod. Toxicol. 13, 491–499. Eldridge, J. C., Stevens, J. T., and Breckenridge, C. B. (2007). Atrazine interaction with estrogen expression systems. Rev. Environ. Contam. Toxicol. 196, 147–160 Review. Eldridge, J. C., and Wetzel, L. T. (2008). Mode of Action of Atrazine for Mammary Tumor Formation in the Female Sprague-Dawley Rat. Chapter 26. In “The Triazine Herbicides” (H.M. Le Baron, McFarland, O. Burnside, and R. Clark, eds.), pp. 399–411. Engel, L. S., Hill, D. A., Hoppin, J. A., Lubin, J. H., Lynch, C. J., Pierce, J., Samanic, C., Sandler, D. P., Blair, A., and Alavanja, M. C. (2005). Pesticide use and breast cancer risk among farmers’ wives in the agricultural health study. Am. J. Epidemiol. 161(2), 121–135. Farber, T. M. (1987). “Pesticide Assessment Guidelines, Subdivision F, Position Document: Selection of a Maximum Tolerated Dose (MTD) in Oncogenicity Studies.” Toxicology Branch, Hazard Evaluation Division, Office of Pesticides Programs, U.S. Environmental Protection Agency, NTIS PB88-116736. Foran, J., and the ILSI Risk Science Working Group on Dose Selection (1997). Principles for the selection of doses in chronic rodent bioassays. Environ. Health Perspect. 105(1), 18–20. French Republic Ministry of Agriculture. (2001). Triazine–based specialties – intended decision. Report by the Toxicity Research Commission on Pesticide Products. Gressel, J., Ammon, H. U., Fogelfors, H., Gasquez, J., Kay, Q. O. N., and Kees, H. (1982). Discovery and distribution of herbicide-resistant weeds outside of North America. In “Herbicide Resistance in Plants,” (H. M. LeBaron and J. Gressel, eds.), pp. 31–46. Wiley, New York. Hauswirth, J. W. and Wetzel, L. T. (1998). Toxicity characteristics of the 2-chlorotriazines, atrazine and simazine. In “Triazine Herbicides: Risk Assessment,” (L. G. Ballantine, J. E. McFarland,

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and D. S. Hackett, eds.), pp. 370–383. American Chemical Society, Washington, DC. Hessel, P. A., Kalmes, R., Smith, T. J., Lau, E., Mink, P. J., and Mandel, J. (2004). A nested case control study of prostate cancer and atrazine exposure. J. Occup. Environ. Med. 46(4), 379–385. Hopenhayn-Rich, C., Stump, M. L., and Browning, S. R. (2002). Regional assessment of atrazine exposure and incidence of breast and ovarian cancer in Kentucky. Arch. Environ. Contam. Toxicol. 42, 127–136. Integrated Risk Information System (IRIS). (1997a). Propazine. Chem. Abst. Service Reg. No. CASRN 139-40-2. USEPA. pp. 1–7. International Agency for Research on Cancer (IARC). (1999). Atrazine. Volume 73. Some Chemicals that Cause Tumours of the Kidney or Urinary Bladder in Rodents and Some Other Substances. p. 59 Jessup, D. C. (1980). “Terbutryn Technical Two-Year Carcinogenicity Study in Mice.” Project Rep. 382–005, International Research and Development Corporation. Mattawan, MI (unpublished). JMPR. (2007). Pesticide residues in food 2007. Joint FAO/WHO Meeting on Pesticide Residues. FAO Plant Production and Protection Paper 191. Kettles, M. A., Browning, S. R., Prince, T. S., and Horstman, S. W. (1997). Triazine herbicide exposure and breast cancer incidence: an ecologic study of Kentucky counties. Environ. Health Perspect. 105(11), 1222–1227. Lebaron, H. M., McFarland, J. E., and Burnside, O. C. (2008). The Triazine Herbicides: A Milestone in the Development of Weed Control Technology. Chapter 1. In “The Triazine Herbicides,” (H. M. Le Baron, McFarland, O. Burnside, and R. Clark, eds.), pp. 1–12. Loosli, R. (1995). Epidemiology of atrazine. Rev. Environ. Contam. Toxicol. 143, 47–57. MacLennan, P., Delzell, E., Sathiakumar, N., Meyers, S. L., Cheng, H., Grizzle, W., Chen, V. W., and Wu, X. C. (2002). Cancer incidence among triazine herbicide manufacturing workers. J. Occup. Environ. Med. 44(11), 1048–1058. Mills, P. K. (1998). Correlation analysis of pesticide use data and cancer incidence rates in California counties. Arch. Environ. Health, 53(6), 410–413. Mills, P. K. and Yang, R. (2003). Prostate cancer risk in California farm workers. J. Occup. Environ. Med. 45(3), 249–258. Muir, K., Rattanamongkogul, S., Smallman-Raynor, M., Thomas, M., Downer, S., and Jenkinson, C. (2004). Breast cancer incidence and its possible spatial association with pesticide application in two counties of England. Public Health 118, 513–520. Neuberger, J. S. (1996). Atrazine and/or triazine herbicides exposure and cancer: An epidemiologic review. J. Agromed. 3(2), 9–30. Osheroff, M. R. (1998). Lifetime oncogenicity study in mice with prometon technical. Hazleton Laboratories Inc. Technical Report. Rusiecki, J. A., De Roos, A., Lee, W. J., Dosemeci, M., Lubin, J. H., Hoppin, J. A., Blair, A., and Alavanja, M. C. (2004). Cancer incidence among pesticide applicators exposed to atrazine in the Agricultural Health Study. J. Natl. Cancer Inst. 96, 1375–1382. Sathiakumur, N., Delzell, E., Austin, H., and Cole, P. (1992). A follow-up study of agricultural chemical production workers. Am. J. Ind. Med. 21, 321–330. Sathiakumur, N. and Delzell, E. (1997). A Review of Epidemiologic Studies of Triazine Herbicides and Cancer. Crit. Rev. Toxicol. 27, 599–612. Sielken, R. L., Valdez-Flores, C., Breckenridge, C., and Stevens, J. (2005). Statistical Inference about the Mechanism of Action in Health, 31, Supplement 1, 151–155. Simpkins, J. W., Eldridge, J. C., and Wetzel, L. T. (1998). Role of strainspecific reproductive patterns in the appearance of mammary tumors

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in atrazine-treated rats Chap. 31. In “Triazine Herbicides: Risk Assessment,” (L. G. Ballantine, J. E. McFarland, and D. S. Hackett, eds.), pp. 399–413. American Chemical Society, Washington, DC. Stevens, J. T., Breckenridge, C. B., Wetzel, L. T., Gillis, J., and Luempert, L. C. III (1994). Hypothesis for mammary tumorigenesis in Sprague– Dawley rats exposed to certain triazine herbicides. J. Toxicol. Environ. Health 43, 139–153. Stevens, J. T., Breckenridge, C. B., Wetzel, L. T., Thakur, A. J., Liu, C., Werner, C., Luempert, L. C. III, and Eldridge, J. C. (1999). A risk characterization for atrazine: Oncogenicity profile. J. Toxicol. Environ. Health 56, 69–109. Sumner, D. D., Luempert, L. C. III, and Stevens, J. T. (1995). Agricultural chemicals: The impact of regulation under FIFRA on science and economics. In “Primer on Regulatory Toxicology,” (C. Chenzelis, J. Holson, and S. Gad, eds.), pp. 133–163. Raven Press, New York. Swindell, R. (2003). Shining light on the evolution of photosynthesis. Answers Magazine 17(3), 74–84. http://images.google.com/ imgres?imgurlhttp://www.answersingenesis.org/tj/images/v17/i3/ photosynthesis_f5.jpg&imgrefurlhttp://www.answersingenesis.org/ tj/v17/i3/photosynthesis.asp&usg__J8IDo252ZPUvKv3tNYm5xC7 5nek&h306&w438&sz33&hlen&start71&um1&tbnid CQYsTlqyH35hhM:&tbnh89&tbnw127&prev/images%3Fq %3Dphotosystem%2BII,%2BD1%2Bprotein%26ndsp%3D20%26hl %3Den%26sa%3DN%26start%3D60%26um%3D1. Thakur, A. J., Wetzel, L. T., Voelker, R. W., and Wakefield, A. E. (1998). Results of a two-year oncogenicity study in Fischer 344 rats with atrazine. In “Triazine Herbicides: Risk Assessment,” (L. G. Ballantine, J. E. McFarland, and D. S. Hackett, eds.), pp. 384–398. American Chemical Society, Washington, DC. Trebst, A. (2008). The Mode of Action of Triazine Herbicides in Plants. In “The Triazine Herbicides,” (H. M. Le Baron, McFarland, O. Burnside, and R. Clark, eds.), pp. 101–118. UK. (2007). Terbuthylazine: Report and proposed decision of the United Kingdom made to the European Commission under Article 8 of 91/414/EEC. Level 2. Appendix 3: List of Endpoints. August draft. UK Rapporteur Monograph. (2000). Council Directive 91/414/EEC. Regulation 3600/92. Atrazine, Volume 3. Annex B. Addendum to the report and proposed decision of the United Kingdom made to the European Commission under Article 7(1) of Regulation 3600/92; Summary, Scientific Evaluation and Assessment. U.S. EPA (1979). “Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA); Good Laboratory Practice Standards—Final Rules. 40 CFR Part 160.” Fed. Reg. 44(91), 27362–27407. U.S. EPA (1982). “Pesticide Assessment Guidelines, Subdivision F. Hazard Evaluation: Human and Domestic Animals.” Environmental Protection Agency-540/9-82-025. Available from NTIS, Springfield, VA. U.S. EPA (1994). Atrazine, simazine, and cyanazine; Notice of initiation of special review. Fed. Reg. 59(225), 60412–60443. U.S. EPA (1995). Terbuthylazine Reregistration Eligibility Decision. EPA 738-R-95-005. http://www.epa.gov/oppsrrd1/REDs/2645.pdf. U.S. EPA (1996). Prometryn Registration Eligibility Decision. EPA-738R-95-033. http://www.epa.gov/oppsrrd1/REDs/0467.pdf. U.S. EPA (1998a). “Pesticide Assessment Guidelines,” http://www.epa. gov/opptsfrs/OPPTS Harmonized/870 Health Effects Test Guidelines/ Series (accessed 3/99). U.S. EPA (2000). Atrazine. Hazard and Dose-Response Assessment and Characterization. SAP Report No. 2000-05. FIFRA Scientific Advisory Panel. June 27-29, 2000. U.S. EPA (2002). Overview of Atrazine Risk Assessment. http://www. epa.gov/oppsrrdI/Reregistration/atrazine/srrdoverview_may02.pdf.

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U.S. EPA (2003a). Atrazine. Interim Reregistration Eligibility Decision (IRED). Docket Number 0062. U.S. EPA (2003b). Revised Atrazine interim Reregistration Eligibility Decision (IRED). Docket ID Number OPP-2003-0367. U.S. EPA (2004). Ametryn: HED Chapter of the Reregistration Eligibility Decision. http://www.epa.gov/oppsrrd1/REDs/ametryn_red.pdf. U.S. EPA (2006a). Cancer Assessment Review Committee, Atrazine: Evaluation of carcinogenic potential. HED DOC. No. 014431. U.S. EPA (2006b). Reregistration Eligibility Decision for Simazine. EPA 738-R-06-008. April 2006. http://www.epa.gov/oppsrrd1/REDs/simazine_red.pdf. Van Leeuwen, J. A., Walmer-Toews, D., Abernathy, T., Smit, B., and Shoukri, M. (1999). Associations between stomach cancer incidence

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and drinking water contamination with atrazine and nitrate in Ontario (Canada) agmecosysterns. 1987-1991. Int. J. Epidem. 28(5), 836–840. Weisburger, J. H. (1975). In “Toxicology, The Basic Science of Poisons,” (L. J. Casarett, and J. Duoll, eds.), MacMillan, New York, p. 333. Wetzel, L. T., Luempert, L. G., Breckenridge, C. B., Tisdell, M. O., Stevens, J. T., Thakur, A. K., Extrom, P. J., and Eldridge, J. C. (1994). Chronic effects of atrazine on estrus and mammary tumor formation in female Sprague-Dawky and Fischer 344 rats. J. Toxicol. Environ. Health 43, 169–182.

Chapter 80

Phenylurea Herbicides Jing Liu Oklahoma State University, Stillwater, Oklahoma

80.1 INTRODUCTION Substituted phenylurea herbicides are a group of pesticides used for general weed control in agricultural and nonagricultural practices – for example, along railroads, utilities’ rightsof-way, and in industrial areas. The first phenylurea herbicide, N,N-dimethyl-N -(4-chlorophenyl)-urea, was introduced in 1952 by DuPont under the common name of monuron. In subsequent years, many more derivatives of this class of compounds have been marketed. The phenylurea herbicides are now manufactured and distributed under the names of anisuron, buturon, chlorbromuron, chlortoluron, chloroxuron, difenoxuron, diuron, fenuron, fluometuron, isoproturon, linuron, methiuron, metobromuron, metoxuron, monuron, neburon, parafluron, siduron, tebuthiuron, tetrafluron, and thidiazuron. Of these, diuron was one of the top 10 most commonly used pesticides in the U.S. industry/commercial/government market sector in 1999 and 2001 (Kiely et al., 2004). The herbicidal action of these compounds is based on their ability to inhibit photosynthesis. Typical phenylurea herbicides are photosystem II inhibitors. Photosystem II is a multisubunit enzyme complex using light energy to catalyze the photooxidation of water to reducing equivalents and oxygen. The reaction center in photosystem II is composed of the proteins D1, D2, CP43, CP47, and the light-harvesting complex II (Rhee et al., 1998). Substituted phenylurea herbicides inhibit photodependent electron transfer by binding to the D1 protein (Arnaud et al., 1994). Degradation of phenylurea herbicides in nature can be a relatively slow process. These pesticides can be decomposed by UV irradiation or by acidic or alkaline conditions. There are four basic types of reactions in the photochemistry of the substituted phenylurea herbicides: photolysis of the C–X bond on aromatics (X  Cl, Br), photoeliminations (Norrish type II reactions), phwotooxidations, and photorearrangements (Kotzias and Korte, 1981). Biological degradation of the compounds in plants and soil is carried out by microflora and microfauna. Vroumsia and coworkers (1996) reported that Rhizoctonia solani Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

(agonomycetes) was the most efficient microorganism tested at degrading diuron, chlortoluron, and isoproturon. The basic reactions of the biotransformation of phenylurea herbicides are N-demethylation followed by oxidation of the aromatic ring. The compounds are gradually transformed by microorganisms to 3-arylureas, which are then metabolized to arylamines, carbon dioxide, and ammonia (Cernakova, 1995; Engelhard et al., 1972). Human cytochrome P-450 3A4 (CYP3A4) expressed in yeast has been reported to catalyze the metabolism of chlortoluron (Mehmood et al., 1995). The metabolism was dependent on NADPH. Chlortoluron was degraded by CYP3A4 into four major metabolites – hydroxylated N-monodemethylated, hydroxylated ring methylated, N-didemethylated, and N-monodemethylated products. Other mammalian cytochrome P-450-mediated reactions have not been reported. Although more than 20 different phenylureas have been marketed for use as herbicides, little information is available on the toxicity of most of these compounds. More specific information on three phenylureas – diuron, fluometuron, and isoproturon – is provided here.

80.2 DIURON Synonyms N ′-(3,4-dichlorophenyl)- N , N -dimethylurea; 3-(3,4-dichlorophenyl)-1,1-dimethylurea; DCMU; DMU Chemical Abstracts Service (CS) number 330541 Molecular formula C9H10Cl2N2O (233.1) Chemical structure Cl

CH3 H N

N

Cl

CH3 O

Trade names and available formulations Karmex, Karmex DL, Diuron 80WP, Diuron 4L, Direx 4L, Di-on, 1725

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Diurex, Duirol, Dailon, Rout, Diater, Unidron, Crisuron, and Cekiuron Physical and chemical properties  Diuron is a white, odorless, crystalline solid with a melting point of 158–159°C and boiling point of 180–190°C. Diuron has a water solubility of approximately 42 ppm (mg/l) at 25°C. At room temperature and neutral pH, hydrolysis of diuron is negligible. Diuron is stable to oxygen and moisture (Worthing, 1983). Usage  Diuron was introduced in 1954 by E. I. du Pont de Nemours & Company under the trademark “Karmex” and is mainly used as a preemergence herbicide for general weed control on noncroplands. It is also used as a soil sterilant in nonagricultural areas (e.g., along railroad rights-of-way). As of 2001, the estimated usage of diuron in the United States was approximately 2–4 million pounds of active ingredients (Kiely et al., 2004). Diuron is also used selectively before emergence on food crop sites, including fruit trees, berries, asparagus, sugarcane, and cotton [U.S. Environmental Protection Agency (EPA), 2003]. In Canada, tolerance for diuron residues was established at 1 ppm on fruits and vegetables (Chapman, 1967). The occupational exposure limit (i.e., 8-h time-weighted average) for diuron in workplace air was established at 10 mg/m3 by the American Conference of Governmental Industrial Hygienists in the United States (International Labor Office, 1980), which indicates that an occupational intake at a rate of approximately 1.4 mg/kg/day is considered safe (Stevens and Sumner, 1991). Absorption, metabolism, and excretion  Diuron is readily absorbed through the gastrointestinal tract in rats and dogs. Tissue levels of diuron were positively correlated with dosage. No apparent storage of diuron in tissues was noted (Hodge et al., 1967). In mammals, diuron is mainly metabolized by dealkylation of the urea methyl groups. Hydrolysis of diuron to 3,4-dichloroaniline and oxidation to 3,4-dichlorophenol as well as hydroxylation at carbon 2 and/or 6 of the benzene ring have also been reported. The predominant metabolite of diuron in urine was N-(3,4-dichlorophenyl)-urea. Diuron is also partially excreted unchanged in feces and urine (Boehme and Ernst, 1965; Hodge et al., 1967). Metabolites found in mammals were qualitatively similar to those found in soil and plants wherein dealkylation was also the major metabolic pathway (Dalton et al., 1966; Geissbuhler et al., 1963). Toxicity to Laboratory Animals Acute toxicity  The oral LD50 (14 days) for diuron in male rats was 3.4 g/kg with 95% confidence limits of 2.9–4.0 g/kg (Hodge et al., 1967). It has been reported (Boyd and Krupa, 1970) that protein content in the diet can influence the acute toxicity of diuron. For example, the LD50 of diuron in weanling rats fed a protein-deficient diet (i.e., approximately 14% of the normal protein intake) was 0.4  0.1 g/kg, whereas the LD50 was 2.4  1.4 g/kg in weanling rats fed a protein-enriched diet. Weanling rats fed

Hayes’ Handbook of Pesticide Toxicology

normal laboratory chow exhibited intermediate sensitivity (LD50  1.0  0.2 g/kg). Signs of acute toxicity following near-lethal dosages of diuron in rats included drowsiness, ataxia, decrease and subsequent increase in reflexes, irritability, and bradypnea. Diarrhea, diuresis, shedding of bloody tears, and nosebleed were also noted. Animals treated with diuron exhibited significant loss of body weight, along with decrease in food and water intake. Hypothermia, glucosuria, proteinuria, and aciduria were detected at 24 h after exposure. Respiratory failure was the immediate cause of death. The intensity of the signs of toxicity was dose dependent. Recovery in surviving animals began at 24 h, and most of the signs disappeared by 72 h after exposure (Boyd and Krupa, 1970). No signs of skin irritation or sensitization were noted in dermally exposed guinea pigs (Hodge et al., 1967). Pathological examination revealed local gastroenteritis, gastric ulcers, and capillary venous congestion of the gastrointestinal mucus. Stress reactions were also seen in the adrenal and thymus glands and in the spleen. Young animals may exhibit signs of kwashiorkor, such as developmental retardation of the adrenal and thymus glands, gastrointestinal tracts, and especially the testes (Boyd and Krupa, 1970). Other pathological changes included an enlarged, congested spleen (Hodge et al., 1967). Subacute toxicity  Rats treated with 1 g/kg of diuron daily for 10 days did not show any lethality but exhibited weight gain retardation. Pathologic changes in the spleen and bone marrow were noted at autopsy 3 and 11 days after the final dose (Hodge et al., 1967). Subchronic and chronic toxicity  Rats fed a diuroncontaining diet (200, 400, 2000, 4000, or 8000 ppm) for 30 days showed growth retardation, and anemia was shown at dosages of 4000 and 8000 ppm. Red blood cell counts and hemoglobin levels were also reduced. Lethality occurred only at 8000 ppm. Congestion and an increase in spleen weight were also observed with the highest dose (Hodge et al., 1967). In 90-day feeding studies (Hodge et al., 1967), 50 and 250 ppm diuron did not cause any toxic effects in rats of either sex. Female rats treated with 500 ppm diuron showed cyanotic discoloration and less weight gain, whereas males exhibited no signs of toxicity. Reductions in red blood cell counts and hemoglobin levels accompanied by a compensatory bone marrow hyperplasia were observed in rats fed 2500 or 5000 ppm diuron. Growth retardation and decreased food consumption were also noted. All these effects of diuron were greater in females than in males. Two-year feeding studies in rats and dogs revealed no significant adverse effects at the dietary levels of 25, 125, or 250 ppm diuron except for an inconsistent and sporadic slight anemia. Growth depression was seen with 2500 ppm diuron (Hodge et al., 1967). Hematotoxicity  Female Sprague–Dawley rats fed 250, 500, or 1000 mg/kg diuron in the diet for 14 months

Chapter | 80  Phenylurea Herbicides

exhibited biochemical and morphological changes in the circulatory system (Wang et al., 1993). Relative spleen weight was significantly increased in a dose-dependent manner. Hemoglobin levels and erythrocyte counts were significantly reduced, whereas methemoglobin concentration and white blood cell counts were increased. Hemoglobin adduct of the released parent aromatic amine, 3,4-DCA, was detected at dose-related levels in animals fed 500 or 1000 mg/kg diuron. Increased pigmentation (hemosiderin) in the spleen was seen histologically, reflecting a response to the hemolytic anemia and methemo­ globinemia induced by the herbicide. Morphological examination of the red blood cells revealed changes such as erythrocytes with the shape of a spindle or with a centrally stained area associated with abnormal hemoglobin, polychromatic erythrocytes, and hypochromic erythrocytes with a large area of central pallor presumably due to the decreased hemoglobin content. Genotoxicity and carcinogenesis  It has been reported that a single dose of 170 or 340 mg/kg diuron given intraperitoneally induced the formation of micronuclei in bone marrow cells in mice at 30 and 48 h after the treatment (Agrawal et al., 1996). Seiler (1978), however, reported that diuron was incapable of inducing micronuclei in erythrocytes when given as a single dose by gavage at 1 or 2 g/kg in mice. In other mutagenicity tests, such as the testicular DNA synthesis inhibition test and the Ames test, diuron exhibited mutagenic activity (Seiler, 1978). Diuron was a suspect genotoxicant, directly or after S-9 activation, at the lowest detected concentrations of 900 or 112.5 g/l, respectively, in the Vibrio fischeri/Mutatox test (CannaMichaelidou and Nicolaou, 1996). Antony and coworkers (1989) reported that topical application of diuron at the rate of 250 mg/kg three times a week for 3 weeks followed by multiple application of a known skin tumor promotor (12-O-tetradecanoyl phorbol 13-acetate) initiated neoplastic transformation and development of skin tumors in mice. Multiple skin applications of diuron alone for up to 52 weeks, however, did not show any tumor-inducing activity. There was no evidence in 2-year bioassays that diuron was carcinogenic in rats or dogs (Hodge et al., 1967). Mice given 464 mg/kg diuron daily from 7 to 28 days of age followed by 1000 ppm diuron daily in the diet for 18 months showed no signs of increased tumor formation (Reinhold, 1987). Teratogenicity  In a study in which the formulation Karmex (containing 80% diuron) was given by gastric intubation to pregnant rats from gestation days 6 to 15 at levels of 125, 250, or 500 mg/kg/day, only the highest dosage reduced both maternal and fetal body weights. Wavy ribs were seen at the dosages of 250 and 500 mg/kg, and delayed ossification of the calvarium was noted in fetuses whose dams received 125 mg/kg diuron (Khera et al., 1979). Diuron showed no teratogenic activity in mice, however (Reinhold, 1987).

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A multigeneration reproductive toxicity study in rats given 125 ppm diuron in the diet revealed no significant changes in reproductive performance endpoints. Postweaning growth of the F2b and F3a generations was moderately affected, however (Hodge et al., 1967). Biochemical effects  Diuron, a dihalogenated substituted urea herbicide, has been reported to be a more potent inducer of hepatic metabolizing enzymes [e.g., benzo(a)pyrene monooxygenase (BP-MOO), 7-ethoxycoumarin O-deethylase (ECOD), and 7-ethoxyresorufin O-deethylase (EROD)] compared to those phenylurea herbicides with one or no halogen substitutions (e.g., chlortoluron and isoproturon) (Schoket and Vincze, 1985, 1986, 1990). Schoket and coworkers (1987) found that repeated diuron exposures (1/6 LD50 for 3 days) decreased the plasma half-life of antipyrine significantly, indicating that hepatic cytochrome P-450 isozymes were induced. Close correlations (r  0.98–0.99) were found between the induction of BP-MOO, ECOD, and EROD and the increase of antipyrine metabolism after diuron treatment. Moreover, hepatic enzymes such as cytochrome P-450, BP-MOO, microsomal epoxide hydrolase, glutathione S-transferase, and UDP-glucuronyltransferase were all induced by diuron in a dose-related manner (oral dosing 1/20 to 1/4 LD50) in rats (Schoket and Vincze, 1990). Dose-related induction of hepatic microsomal enzymes was also seen in rats fed a diuron-containing diet (100, 250, 500, 1000, and 2000 ppm) for 13 weeks (Kinoshita and DuBois, 1970). Maximum induction occurred within the first 3 weeks of feeding and then decreased afterwards. Moreover, a sex difference was noted in the response of the animals; that is, male rats were more sensitive than females to the enzymeinducing activity of diuron. Aquatic toxicity  Diuron used for weed control in water may interfere with the growth of fish and food chain microfauna, such as Daphnia (Crosby and Tucker, 1966). The LC50 of diuron in Daphnia magna or Daphnia pulex at 24 or 48 h is 1.4 mg/l. The LC50s of diuron in the warm water fish Lepomis macrochirus and cold water fish Oncorhynchus kisutch are 7.4 and 16.0 mg/l at 48 h, respectively (Ramamoorthy and Baddaloo, 1995).

80.3  Fluometuron Synonyms  (1,1-dimethyl-3-[3-(trifluoromethyl)phenyl] urea CAS number  2164-17-2 Molecular formula  C10H11F3N2O (232.2) Chemical structure F3C

CH3 H N

N CH3 O

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Trade names and available formulations  Cotoran, Cotoran 4L, Cotoran 85DF, Meturon 4L, Ciba 2059, Cottonex, Lanex, Herbicide C-2059, and Pakhtaran Physical and chemical properties  Fluometuron is a colorless, crystalline sandlike material with a melting point of 163–164.5°C and vapor pressure of 5  107 mmHg at 20°C. Its solubility in water and acetone at 20°C is approximately 105 mg/l (ppm) and 105 g/l (ppt), respectively. Fluometuron is soluble in most organic solvents [International Agency for Research on Cancer (IARC), 1983; Worthing, 1983]. Usage  Fluometuron has been used as a herbicide in the United States for more than three decades. Since being introduced as a commercial chemical in 1960 by CibaGeigy AG under the trademark Cotoran (Worthing, 1983), fluometuron has been widely used to control broadleaf weeds and grasses on agricultural crops (e.g., cotton and sugarcane). The amount of fluometuron used in the United States in 1976 and 1978 was estimated to be 5.3 and 2.9 million pounds, respectively (IARC, 1983). As of 1995, the approximate quantity of fluometuron used annually in U.S. agricultural practices was approximately 5–9 million pounds (Aspellin, 1997). Fluometuron is currently registered by the U.S. EPA for use on cotton only. An estimated 2.4 million pounds of fluometuron was used annually in the United States from 1998 to 2002 (U.S. EPA, 2005). The tolerance for fluometuron in or on raw agricultural commodities, cottonseed, and sugarcane was set at 0.1 ppm in the United States (U.S. EPA, 1980a). In or on sugarcane bagasse, a tolerance of 0.2 ppm was established (U.S. EPA, 1980b). Maximum occupational exposure to fluometuron was established at the level of 5 mg/m3 in workplace air in the former Soviet Union (International Labor Office, 1980). Toxicity to Laboratory Animals Acute toxicity  The oral LD50 of fluometuron was approximately 8.9 g/kg in rats of both sexes (Ben-Dyke et al., 1970), 0.9 and 2.4 g/kg in male and female mice, and greater than 10 g/kg in rabbits and dogs (Spencer, 1968). The dermal LD50 in rats and rabbits was greater than 2 and 10 g/kg, respectively (Ben-Dyke et al., 1970; Worthing, 1983). LC50s (96 h) for rainbow trout, crucian carp, and bluegill are 47, 170, and 96 mg/l, respectively (Worthing, 1983). Animals treated with lethal dosages of fluometuron exhibited signs of depression, gasping, hyperpnea, lacrimation, and peripheral vasoconstriction (National Cancer Institute, 1980). Toxicity of fluometuron in 6- to 9-month-old desert sheep has been seen with a single oral dose of 0.8 or 4 g/kg (Mohamed et al., 1995). Signs of toxicity appeared within 15 min after exposure and included depression, salivation, grinding of the teeth, chewing movement of the jaws, mydriasis, dyspnea, incoordination of movements, and drowsiness. Similar signs of toxicity were also seen in animals treated with repeated daily dosages of 25 or 200 mg/kg

fluometuron. Laboratory testing revealed increased activities of serum alanine aminotransferase (ALT), aspartate transaminase (AST), and lactate dehydrogenase. Blood urea nitrogen was also elevated. Total serum protein and calcium were significantly decreased. Subchronic toxicity  In 90-day feeding studies (National Cancer Institute, 1980) in which rats or mice of both sexes were treated with 250, 500, 1000, 2000, 4000, 8000, or 16,000 ppm fluometuron, less weight gain was seen with the three highest doses in both male and female rats. Deaths occurred in male rats fed 8000 and 16,000 ppm fluometuron and in females that received 16,000 ppm. Various degrees of spleen enlargement were observed in rats of both sexes treated with 2000 ppm or more of fluometuron. Dose-related pathological changes in rats included mild to severe congestion of the red pulp with corresponding atrophy of the white pulp and depletion of the lymphocytic elements. There were essentially no signs of toxicity observed in mice in these subchronic studies except for a moderate decrease (10%) in body weight gain in both sexes at levels of 4000 ppm and greater. Mutagenicity  Fluometuron given as a single dose by gavage in mice exhibited mutagenic activity in both the testicular DNA synthesis inhibition test and the erythrocyte micronucleus test at levels of 1 and 2 g/kg. In the in vitro Ames test, fluometuron also showed mutagenic activity (Seiler, 1978). Carcinogenicity  Mice of both sexes (7 weeks old) fed a diet containing �������������������������������������������� 500 or 1000 mg/kg fluometuron for 103 weeks showed similar incidences of both neoplastic and nonneoplastic lesions compared to the control animals. A nonsignificant increase in the incidences of hepatocellular adenomas or carcinomas and tumors in the hematopoietic system (e.g., lymphoma and leukemia) was seen in male mice. Carcinogenicity studies in rats fed a diet ���������������� containing 125 or 250 mg/kg fluometuron for 103 weeks were negative (IARC, 1983; National Cancer Institute, 1980).

80.4  Isoproturon Synonyms  N,N-dimethyl-N′-[4-(1-methylethyl)phenyl] urea; 3-(4-isopropylphenyl)-1,1-dimethylurea CAS number  34123-59-6 Molecular formula  C12H18N2O (206.3) Chemical structure CH3 H3C

H N

N CH3

H3C

O

Trade names and available formulations  Arelon, Graminon

Chapter | 80  Phenylurea Herbicides

Physical and chemical properties  Isoproturon is a colorless powder with a melting point of 155–156°C and vapor pressure of 2.5  108 mmHg at 20°C. Its water solubility is approximately 55 ppm at 20°C. Isoproturon is soluble in most organic solvents and stable to light, acids, and alkali (Worthing, 1983). Usage  Isoproturon was introduced as a herbicide by Hoechst AG under the trade name Arelon, by Ciba-Geigy AG under the trade name Graminon, and by Rhone-Poulenc as phytosanitaire. Isoproturon is used to control selectively germinating broadleaf and grass weeds in crops such as wheat, cereal, sugarcane, citrus, cotton, and asparagus. It is widely used in many countries. Isoproturon is registered for use in countries such as India and European Union members (Pesticide Action Network pesticide database – ­international registration information for isoproturon available at www. pesticideinfo.org). Lebailly et al. (2009) reported that in France isoproturon is the most heavily used herbicide on wheat and barley. Toxicity to Laboratory Animals Basic findings  The acute oral LD50 of isoproturon in rats was estimated at 1.8–2.4 g/kg, whereas the acute dermal LD50 was greater than 3.2 g/kg (Worthing, 1983). An acute oral LD50 of 1 g/kg in rats has also been reported (Behera and Bhunya, 1990). In 90-day feeding studies, the no-observed-adverseeffect level was reported to be 400 mg/kg/day in rats and 50 mg/kg/day in dogs (Worthing, 1983). The 96-h LC50s of isoproturon for carp and rainbow trout were 193 and 240 mg/l, respectively (Worthing, 1983). Subacute and subchronic toxicity  Repeated dermal exposure of isoproturon technical (IPT; 250, 500, or 1000 mg/kg/day) and its wettable powder formulation (IPF; 750, 1500, or 2250 mg/kg/day) in rats of both sexes for 21 days produced no clear overt signs of toxicity (Dikshith et al., 1990). Increased organ:body weight ratios (e.g., liver, kidney, adrenal, and spleen) were seen more in the females with both IPT and IPF, especially with the highest dosages. Hematological studies showed that all three dosages of IPT decreased red blood cell counts in males, whereas only the highest dosage caused a slight reduction in erythrocyte counts in females. Hemoglobin levels in both sexes were reduced by all three dosages of isoproturon (IPT). Neutrophils were decreased and lymphocytes were increased by all IPT exposures. IPF, on the other hand, did not produce any hematological changes. Activities of ALT and AST and protein content in the liver and serum were also altered by IPT or IPF in one or both sexes of the rats. For example, females had an increase in serum AST activity, a decrease in liver ALT, and a significant reduction in serum protein after repeated IPT. Repeated IPF exposure reduced serum ALT activity in male rats and lowered total serum proteins in female rats.

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No overt signs of toxicity were observed in rats treated orally with 200, 400, or 800 mg/kg/day of isoproturon for 42 and 60 days (Sarkar et al., 1995). The highest dosage of the chemical, however, did significantly decrease body weight. Moreover, a dose-dependent increase in the liver weight was noted. Isoproturon also increased the weights of the kidney and the heart. Histopathological changes included hepatocellular degeneration and focal necrosis in the liver, glomerular and tubular degeneration in the kidney, and hemosiderosis in the spleen. Genotoxicity  Isoproturon and its structural analogs fluometuron and monuron have all been reported to be genotoxic (Behera and Bhunya, 1990; Garrett et al., 1986; Seiler, 1978). Behera and Bhunya (1990) found that 100, 150, and 200 mg/kg of isoproturon given intraperitoneally to adult Swiss albino mice induced various types of chromosomal aberrations in bone marrow cells, such as chromatid gaps, chromatid breaks, acentric fragments, chromatid exchanges, ring chromosomes, and metacentric chromosomes. The highest dosage of isoproturon also induced the formation of micronuclei in bone marrow cells. Pregnant rats treated orally with 180 mg/kg of isoproturon daily from gestation day 6 to day 20 also exhibited chromatid breaks in bone marrow cells (Srivastava and Raizada, 1995). Reproductive toxicity  Isoproturon has been found to induce abnormalities in sperm shape in a dose-dependent manner in mice (Behera and Bhunya, 1990). The sperm would have either hammer-, mushroom-, or amorphousshaped heads or hook- and beak-shaped acrosomal ends. Sarkar and co-workers (1997) also reported the potential toxic effects of isoproturon on the male reproductive system in rats. When isoproturon was given orally to rats 6 days/ week at a rate of 200, 400, or 800 mg/kg/day for 10 weeks, the highest dosage decreased epididymal sperm counts and the percentage of motile sperm with an increased percentage of abnormal sperm (e.g., distorted heads, atypical tails, and bent necks or midpieces). Degeneration and desquamation of germinal layer cells were observed in the testis. Tubular lumens of the testis and the epididymis exhibited a reduced number of spermatids and spermatozoa, respectively (Sarkar et al., 1995). The activities of androgen biosynthesis-related enzymes (e.g., glucose-6-phosphate dehydrogenase and 5-3--hydroxysteroid dehydrogenase) were reduced in a dose-related manner (Sarkar et al., 1997). Overall, isoproturon has been suggested to have the potential to cause maturational malformation of sperm cells and retarded spermatogenesis in rats. Fetotoxicity and teratogenicity  Isoproturon (45, 90, or 180 mg/kg/day) given orally to pregnant rats from gestation day 6 to day 20 caused no observable fetotoxic and/ or teratogenic effects. The numbers of implantations and resorptions, fetal body weights, and external, visceral, and skeletal structures were all comparable to those of the controls (Srivastava and Raizada, 1995). However, with higher

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dosages of isoproturon (225, 450, and 900 mg/kg/day) given orally from gestation day 6 to day 15 (Sarkar and Gupta, 1993a,b), dose-related depression and drowsiness of the dams were observed. Although there was no lethality associated with isoproturon exposures, decreased maternal body weight was noted during later pregnancy (gestation days 15–20) with the dosages of 450 and 900 mg/kg. Litter size, fetal weights, and crown–rump and transumbilical lengths were all decreased by isoproturon dosages of 450 mg/kg or more. Moreover, there was a significant increase in the frequency of fetal resorptions and the number of fetuses with retarded growth. Again, no major visceral and/or skeletal malformations were observed. Neurological effects Neurotoxic effects of isoproturon in mice have been reported by Sarkar and Gupta (1993a,b). They found that a single oral dose of isoproturon (0.5, 1.0, or 2.0 g/kg) potentiated both pentobarbital- and barbitalinduced sleeping time. Spontaneous and forced locomotor activities were reduced by 1.0 and 2.0 mg/kg isoproturon. In addition, isoproturon exhibited anticonvulsant activity against electroshock and pentylenetetrazol-induced convulsions. The mechanisms of these inhibitory effects of isoproturon on the central nervous system are unclear. Biochemical effects Isoproturon given by gavage to rats at the dose of 1/6 LD50 for 3 consecutive days induced significantly the activities of hepatic enzymes (e.g., NADPHcytochrome c reductase, microsomal epoxide hydrolase, 7-ethoxycoumarin O-deethylase, aldrin epoxidase, UDPglucuronyl-transferase, and glutathione S-transferase) (Schoket and Vincze, 1985, 1986). Antipyrine plasma halflife, however, was not affected by the isoproturon treatment, suggesting that hepatic cytochrome P-450 isozymes [e.g., benzo(a)pyrene monooxygenase and 7-ethoxyresorufin Odeethylase] were not induced by isoproturon (Schoket et al., 1987). Hazarika and Sarkar (2001) reported that isoproturon, when given orally (675 mg/kg/day for 3 days) in rats, significantly increased lipid peroxidation in brain and liver while glutathione levels and glutathione-S-transferase activity were significantly increased. A 90-day study with oral administration of technical isoproturon (22.5, 45, and 90 mg/kg/day) in rats revealed few apparent signs of toxicity except for changes in liver alkaline phosphatase and lactate dehydrogenase activities and a significant decrease in the white blood cell count with higher dosages (Raizada et al., 2001).

CONCLUSION The phenylurea pesticides have been used for decades as preemergent herbicides. Although approximately 20 different phenylureas have been marketed, there is relatively little toxicity information available for most. Diuron, the most common phenylurea, has been among the top 10 pesticides in use in the United States. The acute toxicity potential for all phenylureas appears to be low, with oral LD50 values typically greater than 1 g/kg. With high-level

acute or repeated exposures, neurobehavioral alterations, body weight reductions, hematotoxicity, and hepatotoxicity have been reported.

REFERENCES Agrawal, R. C., Kumar, S., and Mehrotra, N. K. (1996). Micronucleus induction by diuron in mouse bone marrow. Toxicol. Lett. 89, 1–4. Antony, M., Shukla, Y., and Mehrotra, N. K. (1989). Tumor initiatory activity of a herbicide diuron on mouse skin. Cancer Lett. 48, 125–128. Arnaud, L., Taillandier, G., Kaouadji, M., Ravanel, P., and Tissut, M. (1994). Photosynthesis inhibition by phenylureas: A QSAR approach. Ecotoxicol. Environ. Safety 28, 121–133. Behera, B. C., and Bhunya, S. P. (1990). Genotoxic effect of isoproturon (herbicide) as revealed by three mammalian in vivo mutagenic bioassays. Ind. J. Exp. Biol. 28, 862–867. Ben-Dyke, R., Sanderson, D. M., and Noakes, D. N. (1970). Acute toxicity data for pesticides. World Rev. Pest. Control 9(3), 119–127. Boehme, C., and Ernst, W. (1965). The metabolism of urea-herbicides in the rat: 2. Diuron and linuron. Food Cosmet. Toxicol. 3, 797–802 [in German]. Boyd, E. M., and Krupa, V. (1970). Protein-deficient diet and diuron toxicity. J. Agric. Food Chem. 18(6), 1104–1107. Canna-Michaelidou, S., and Nicolaou, A.-S. (1996). Evaluation of the genotoxicity potential (by Mutatox™ test) of ten pesticides found as water pollutants in Cyprus. Sci. Total Environ. 193, 27–35. Cernakova, M. (1995). Biological degradation of isoproturon, chlortoluron and fenitrothion. Folia Microbiol. 40(2), 201–206. Chapman, R. A. (1967). “Tolerances for Residues of Pesticide Chemicals,” T.I.L. No. 290. Food and Drug Directorate, Department of National Health and Welfare, Ottawa, Ontario, Canada. Crosby, D. G., and Tucker, R. K. (1966). Toxicity of aquatic herbicides to Daphnia magan. Science 154, 289–290. Dalton, R. L., Evans, A. W., and Rhodes, R. C. (1966). Disappearance of diuron in cotton field soils. Weeds 14, 31–33. Dikshith, T. S. S., Raizada, R. B., and Srivastava, M. K. (1990). Dermal toxicity to rats of isoproturon technical and formulation. Vet. Hum. Toxicol. 32(5), 432–434. Engelhard, G., Wallnofer, P. R., and Plapp, K. (1972). Identification of N,O-dimethylhydroxylamine as microbial degradation product of the herbicide linuron. Appl. Microbiol. 23, 664–666. Garrett, N. E., Stack, H. F., and Waters, M. D. (1986). Evaluation of the genetic activity profiles of 65 pesticides. Mutat. Res. 168, 301–325. Geissbuhler, H. C., Haselback, C., Aebi, H., and Ebner, L. (1963). The fate of N -(4-chlorophenoxy)-phenyl-N,N-dimethylurea (C-1983) in soils and plants: III. Breakdown in soils. Weed Res. 3, 277–297. Hazarika, A., and Sarkar, S. N. (2001). Effect of isoproturon pretreatment on the biochemical toxicodynamics of anilofos in male rats. Toxicology 165(2-3), 87–95. Hodge, H. C., Downs, W. L., Panner, B. S., Smith, D. W., and Maynard, E. A. (1967). Oral toxicity and metabolism of diuron (N-(3,4-dichlorophenyl)-N’,N -dimethylurea) in rats and dogs. Food Cosmet. Toxicol. 5, 513–531. International Agency for Research on Cancer (IARC) (1983). Fluometuron. In “IARC Monographs on the Evaluation of the Carcinogenic Risk of Chemicals to Man,” pp. 245–253. IARC, Lyon, France. International Labor Office (1980). “Occupational Exposure Limits for Airborne Toxic Substances,” Occupational Safety and Health Series No. 37, 2nd rev. ed., pp. 106, 118–119. International Labor Office, Geneva.

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Khera, K. S., Whalen, C., Trivett, G., and Angers, G. (1979). Teratogenicity studies on pesticidal formulations of dimethoate, diuron and lindane in rats. Bull. Environ. Contam. Toxicol. 22, 522–529. Kiely, T., Donaldson, D., and Grube, A. (2004). “Pesticides Industry Sales and Usage: 2000 and 2001 Market Estimates.” Biological and Economic Analysis Division, Office of Pesticide Programs, Office of Prevention, Pesticides and Toxic Substances Available at www.epa. gov. U.S. Environmental Protection Agency, Washington, DC. Kinoshita, F. K., and DuBois, K. P. (1970). Induction of hepatic microsomal enzymes by herban, diuron and other substituted urea herbicides. Toxicol. Appl. Pharmacol. 17, 406–417. Kotzias, D., and Korte, F. (1981). Photochemistry of phenylurea herbicides and their reactions in the environment. Ecotoxicol. Environ. Safety 5, 503–512. Lebailly, P., Bouchart, V., Baldi, I., Lecluse, Y., Heutte, N., Gislard, A., and Malas, J. P. (2009). Exposure to pesticides in open-field farming in France. Ann. Occup. Hyg. 53, 69–81. Mehmood, Z., Kelly, D. E., and Kelly, S. L. (1995). Metabolism of the herbicide chlortoluron by human cytochrome P450 3A4. Chemosphere 31(11/12), 4515–4529. Mohamed, O. S. A., Ahmed, K. E., Adam, S. E. I., and Idris, O. F. (1995). Toxicity of cotoran (fluometuron) in desert sheep. Vet. Hum. Toxicol. 37(3), 214–216. National Cancer Institute (1980). “Bioassay of Fluometuron for Possible Carcinogenicity.” Carcinogenesis Technical Report Series. CAS No. 2164-17-2, NCI-CG-TR-195, NTP-80-11. Raizada, R. B., Srivastava, M. K., Kaushal, R. A., Singh, R. P., and Gupta, K. P. (2001). Subchronic oral toxicity of a combination of insecticide (HCH) and herbicide (ISP) in male rats. J. Appl. Toxicol. 21(1), 75–79. Ramamoorthy, S., and Baddaloo, E. G. (1995). Aquatic toxicity data, pp. 165, 172, 251, 284. In “Handbook of Chemical Toxicity Profiles of Biological Species: Aquatic Species.” CRC Press, Boca Raton, FL. Reinhold, V. N. (1987). Diuron, a review. Danger. Prop. Indust. Mater. Rep. 7(5), 49–55. Rhee, K. H., Morris, E. P., Barber, J., and Kuhlbrandt, W. (1998). Threedimensional structure of the plant photosystem II reaction centre at 8 Å resolution. Nature 398(19), 283–286. Sarkar, S. N., and Gupta, P. K. (1993a). Fetotoxic and teratogenic potential of substituted phenylurea herbicide, isoproturon, in rats. Ind. J. Exp. Biol. 31, 280–282. Sarkar, S. N., and Gupta, P. K. (1993b). Neurotoxicity of isoproturon, a substituted phenylurea herbicide, in mice. Ind. J. Exp. Biol. 31, 977–981. Sarkar, S. N., Chattopadhyay, S. K., and Majumdar, A. C. (1995). Subacute toxicity of urea herbicide, isoproturon, in rats. Ind. J. Exp. Biol. 33, 851–856.

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Sarkar, S. N., Majumdar, A. C., and Chattopadhyay, S. K. (1997). Effect of isoproturon on male reproductive system: Clinical, histological and histoenzymological studies in rats. Ind. J. Exp. Biol. 35, 133–138. Schoket, B., and Vincze, I. (1985). Induction of rat hepatic drug metabolizing enzymes by substituted urea herbicides. Acta Pharmacol. Toxicol. 56, 283–288. Schoket, B., and Vincze, I. (1986). Induction of rat hepatic microsomal epoxide hydrolase by substituted urea herbicides. Acta Pharmacol. Toxicol. 58, 156–158. Schoket, B., and Vincze, I. (1990). Dose-related induction of rat hepatic drug-metabolizing enzymes by diuron and chlorotoluron, two substituted phenylurea herbicides. Toxicol. Lett. 50, 1–7. Schoket, B., Zilahy, Z., Molnar, J., and Vincze, I. (1987). Comparative investigation of antipyrine half-life and induction of cytochrome P-450 dependent monooxygenases in rats treated with phenylurea herbicides. In Vivo 1, 185–188. Seiler, J. P. (1978). Herbicidal phenylalkylureas as possible mutagenicity tests with some urea herbicides. Mutat. Res. 58, 353–359. Spencer, E. Y. (1968). “Guide to the Chemicals Used in Crop Protection.” Research Branch Agriculture Canada Publication No. 1093 5th ed. University of Western Ontario, London, Ontario, Canada. Srivastava, M. K., and Raizada, R. B. (1995). Developmental toxicity of the substituted phenylurea herbicide isoproturon in rats. Vet. Hum. Toxicol. 37(3), 220–223. Stevens, J. T., and Sumner, D. D. (1991). Herbicides. In “Handbook of Pesticide Toxicology”: Vol. 3. Classes of Pesticides (W. J. Hayes and E. R. Laws, eds.), pp. 1349–1350. Academic Press, San Diego. U.S. Environmental Protection Agency (1980a). “Protection of Environment.” US Code Fed. Regul. Title 40, part 180.229. U.S. Environmental Protection Agency (1980b). “Food and Drugs.” US Code Fed. Regul., Title 21, part 561.240. U.S. Environmental Protection Agency (2003, September). “Registration Eligibility Decision (RED) for Diuron.” Available at www.epa. gov/pesticides. U.S. Environmental Protection Agency (2005, September). “Registration Eligibility Decision (RED) for Fluometuron.” Available at www.epa. gov/pesticides. Vroumsia, T., Steiman, R., Seigle-Murandi, F., Benoit-Guyod, J. L., and Khadrani, A. (1996). Biodegradation of three substituted phenylurea herbicides (chlortoluron, diuron, and isoproturon) by soil fungi. A comparative study. Chemosphere 33(10), 2045–2056. Wang, S. W., Chu, C. Y., Hsu, J. D., and Wang, C. J. (1993). Haemotoxic effect of phenylurea herbicides in rats: Role of haemoglobin-adduct formation in splenic toxicity. Food Chem. Toxicol. 31(4), 285–295. Worthing, C. R. (1983). “The Pesticide Manual: A World Compendium,” pp. 226, 281, 329 7th ed. British Crop Protection Council, Hampshire, UK.

Chapter 81

Protoporphyrinogen Oxidase-Inhibiting Herbicides Franck E. Dayan and Stephen O. Duke United States Department of Agriculture, University, Mississippi

81.1  Introduction Protoporphyrinogen oxidase-inhibiting herbicides (also referred to as Protox- or PPO-inhibiting herbicides) were commercialized in the 1960s and their market share reached approximately 10% (total herbicide active ingredient output) in the late 1990s. The widespread adoption of glyphosate-resistant crops (Duke and Powles, 2008) has caused a significant reduction of the field application of Protox inhibitors, and these herbicides accounted for only 1.3% of the total herbicide output in the United States in 2006 [U.S. Department of Agriculture, National Agricultural Statistics Service (USDA–NASS), 2008]. Before the molecular target site of these herbicides was discovered (Matringe et al., 1989a,b), these compounds were often termed “diphenyl ether–type herbicides.” At that time, almost all of the Protox inhibitors were diphenyl ethers. This nomenclature led to some confusion in herbicide classification because other “diphenyl ether” herbicides have an entirely different molecular site of action (i.e., inhibition of acetyl CoA carboxylase). Now, many other structural classes of Protox inhibitors are commercialized. In general, the newer products are more potent Protox inhibitors, resulting in lower application rates than the older herbicides of this class. Some of them appear to be analogs of the substrate or a substrate/product transition state of the enzyme (Reddy et al., 1998). There are several previous reviews of the Protox inhibitors, both before (Böger, 1984; Gilham and Dodge, 1987; Kunert et al., 1987; Matsunaka, 1976) and after (Dayan and Duke, 1996, 1997; Duke et al., 1991; Nandihalli and Duke, 1993; Reddy et al., 1998; Scalla and Matringe, 1994) their target site was known. The emphasis in most of these previous reviews was on the mode of action. An entire book dealing mainly with Protox inhibitors is available (Duke and Rebeiz, 1994). This review updates and supplements a chapter published in the second edition of the Handbook of Pesticide Toxicology (Dayan et al., 2001). Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

81.2  Commercially available protox inhibitors 81.2.1  Diphenyl Ether Protoporphyrinogen Oxidase Inhibitors Nitrofen was the first Protox-inhibiting herbicide to be introduced for commercial use in 1964. This diphenyl ether (DPE) herbicide was eventually recognized as a relatively weak inhibitor of Protox (Nandihalli et al., 1992), but it was a lead compound of an entire class of structurally related herbicides that were much more active. Several DPE herbicides (Table 81.1) have been successfully commercialized (Anderson et al., 1994). Although many of these older commercialized DPEs have a p-nitrophenyl substitution, newer DPE-like herbicides more commonly contain p-trifluoromethyl phenyl substitutions. These new herbicides are heterocyclic phenyl ethers (structurally related to DPEs), where one of the phenyl rings of ether is replaced by an aromatic heterocycle. Reported examples include 6-aryloxy-1H-benz­ otriazoles (Condon et al., 1995), aryloxyindolin-2(3H)ones (Karp et al., 1995), 5-aryloxybenzisoxazoles (Wepplo et al., 1995), 6-aryloxyquinoxalin-2,3-diones (Anderson et al., 1994), benzheterocycles (Lee et al., 1995), and benzoxazines (Sumida et al., 1995). Pyrazolyl, pyridyl, and furyl rings have also been investigated as the heterocyclic component (Anderson et al., 1994; Armbruster et al., 1993; Clark, 1994, 1996; Sherman et al., 1991).

81.2.2  Non–Diphenyl Ether Protoporphyrinogen Oxidase Inhibitors Whereas the first generation of Protox inhibitors (with the exception of oxadiazon) was primarily based on the DPE backbone, numerous nonoxygen-bridged compounds with this same site of action have been discovered (Table 81.2). These compounds invariably consist of heterobicyclic 1733

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Table 81.1  Commercially Available DPE and DPE-like Protox-Inhibiting Herbicidesa Common name

Trade name

Primary source

Acifluorfen

Blazer ®

BASF

Chemical structure Cl

COOH

F3C

Bifenox

Foxpro� ®

Bayer Crop Science

O

Cl

COOCH3

Cl

Fluoroglycofen

Compete� ®

Dow AgroSciences

O

Reflex� ®

COOCH2COOH

Syngenta Crop Protection

O

Cobra� ®

NO2

O Cl

Valent

O

C NH S

F3C

Lactofen

NO2

Cl F3C

Fomesafen

NO2

O

NO2

O

NO2

CH3

O

Cl F3C

COOCH(CH3)COOCH2CH3

Oxyfluorfen

Goal� ®

Dow AgroSciences

Cl F3C

OCH2CH3 O

NO2

a

Information from Senseman (2007).

structures with one phenyl ring attached to a heterocyclic ring. The linkage can consist of a carbon–carbon bridge, as with the isoxazole carboxamides (Dayan et al., 1997a; Hamper et al., 1995), but more often consists of a carbon– nitrogen bridge, as with the phenyl imides (Huang et al., 2005; Lyga et al., 1991; Mito et al., 1991; Sato et al., 1987), triazolinones (Amuti et al., 1997; Dayan et al., 1997b,c; Luo et al., 2008; Theodoridis, 1997; Theodoridis et al., 1992, 1995), oxadia-zolones (Dickmann et al., 1997), and pyrazoles (Prosch et al., 1997) and uracils (Hou et al., 2005; Zhang et al., 2006). Before glyphosate-resistant crops, there were several reasons for the high level of interest among agrochemical companies in the discovery and development of new Protox inhibitors. First, weeds have not shown a propensity to evolve resistance to these herbicides (see Section 81.4.3). In addition, the market niche for Protox inhibitors was beginning to expand to weed control in monocot crops, with the marketing of newer compounds such as carfentrazone, fluazolate, and oxadiargyl (Table 81.3). Finally, Protox is a particularly good target that can be inhibited by structurally diverse classes of herbicides, allowing for the development of a new chemistry not yet patented by other companies.

81.3  Agricultural use 81.3.1  Crops and Weeds Protox inhibitors have historically been used for broadspectrum weed control in soybean fields (see Table 81.3). However, because Protox-inhibiting herbicides control both monocotyledonous and dicotyledonous weeds, Dayan and Duke (1996) proposed that the market share of Protox inhibitors could be increased by developing compounds for use in monocot crops. Several non-DPE herbicides based on the triazolinone and the oxadiazole structures have been, or will soon be, registered for weed control in cereal crops. Currently, most of the compounds with selectivity for cereals and small grains (bifenox, carfentrazone, and fluoroglycofen) are not available in the U.S. market. Finally, compounds with the highest biological activity, such as oxyfluorfen and azafenidin, have also been developed for use as nonselective herbicides in noncrop areas and nurseries. In approximately the past decade, glyphosate used with glyphosate-resistant soybeans has come to dominate weed management in soybeans in the Americas, resulting in greatly diminished use of other herbicides (Duke and

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Table 81.2  Nonoxygen-Bridged Protox-Inhibiting Herbicidesa Common name

Trade name

Primary source

Azafenidin

Evoluso� ®

DuPont

Chemical structure O

OCH2C

N

Cl

CH

N N Cl

Butafenacil

®

Inspire� 

Syngenta Crop Protection

O N Cl

N

O

CF3

O O O O

Carfentrazone-ethyl

Addit� ®

FMC

O

F

CHF2 N

Cl

N N

CH3

CH2CHClCOOCH2CH3

Fluazolate

—

Bayer Crop science

F Br

CF3

Cl N

CH3

O O

Flufenpyr-ethyl

—

CH(CH3)2

Valent

F

O

Cl

N

CF3 N

O

O

O

Flumiclorac

®

Resource� 

Valent O

OCH2COOH

N

Cl

O

Flumioxazin

®

Valor� 

Valent

F

O

F

N

O

O N O

(Continued)

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Table 81.2  (Continued) Common name Fluthiacet-methyl

Trade name ®

Appeal� 

Primary source

Chemical structure

K-I Chemical USA

F

N

N N

Cl

S

O

S

O O

Oxadiargyl

Raft� ®

Bayer Crop Science Cl

O O

Cl

N N

C(CH3)3

O CH2C

Oxadiazon

Ronstar� ®

CH

Bayer Crop Science (CH3)2HC

O

C(CH3)3

N N

Cl

O

Cl

Pyraflufen-ethyl

Milan� ®

Nichino America F N Cl O

O

N O

Cl

CHF2

O

Sulfentrazone

®

Authority� 

FMC Corporation

Cl O

CHF2 N

Cl

N N

CH3

NHSO2CH3 a

Information from Senseman (2007).

Powles, 2008; Gianessi, 2008). However, both natural and evolved glyphosate-resistant weeds are becoming a problem that must be addressed to some extent with other herbicides (Duke and Powles, 2008; Nandula et al., 2005). Protox inhibitor herbicides could be increasingly useful tools in dealing with this problem.

81.3.2  Mode of Application For the most part, Protox-inhibiting herbicides are applied postemergence during the early stages of weed development. The rate of application ranges widely among Protox

inhibitors. Most of the DPE herbicides (e.g., acifluorfen and lactofen) are applied at rates of 100 to 500 g active ingredient (a.i.)/ha. Fluoroglycofen is more active and can be applied at rates 10-fold lower. The most active of these herbicides can be applied at rates as low as 1 g a.i./ha. With a few exceptions, these compounds do not have preemergence activity and have little residual activity in soils. However, some structures, such as sulfentrazone, have excellent preemergence activity (see Table 81.3) (Theodoridis et al., 1992). The preemergence-applied Protox inhibitors are active at concentrations ranging from 2 to 4 kg a.i./ha for oxadiazon to 100 g a.i./ha for sulfentrazone.

Chapter | 81  Protoporphyrinogen Oxidase-Inhibiting Herbicides

81.4  Behavior in plants

Table 81.3  Targeted Crops and Mode of Application of Protox-Inhibiting Herbicidesa

Protox inhibitors cause rapid photobleaching and lightdependent desiccation of foliage. The symptoms observed on the foliage of DPE-treated plants include leaf cupping, crinkling, bronzing, and necrosis (Johnson et al., 1978). The lesions on the foliage are due to loss of membrane integrity that leads to cellular leakage (Figure 81.1) (Dayan et al., 1997c, 1998; Kenyon et al., 1985; Lee et al., 1995). Other physiological responses include inhibition of photo­ synthesis; evolution of ethylene, ethane, and malondi­ aldehyde; and, finally, bleaching of chloroplast pigments (Kenyon et al., 1985). Protox inhibitors are known to cause temporary injury to the foliage of treated crops (particularly soybean) (Graham, 2005). However, crops generally recover rapidly, and yields are not affected (Vidrine et al., 1993, 1994, 1996; Walker et al., 1992). Although not desirable, crop injury on soybean is not unusual and farmers have become accustomed to this phenomenon with a number of herbicide classes. Also, treatment with Protox inhibitors can induce the expression of pathogen defenserelated genes and the production of phytoalexins, which can enhance plant resistance to some diseases (Dann et al., 1999; Graham, 2005; Komives and Cassida, 1982, 1983; Sanyal and Shrestha, 2008; Smith and Hallett, 2006). In fact, the commercial label of the Protox-inhibiting herbicide lactofen lists white mold suppression in soybeans as one of its properties (Duke et al., 2007).

a

Common name

Main crop

Application

Acifluorfen

Soybean, peanut, rice

Postemergence

Azafenidin

Perennial crops/ forestry

Preemergence

Bifenox

Small grain

Preemergence/ postemergence

Butafenacil

Cotton defoliant

Postemergence

Carfentrazone-ethyl

Cereal crops

Postemergence

Fluazolate

Winter wheat

Preemergence

Flufenpyr-ethyl

Soybean

Postemergence

Flumiclorac

Soybean and maize

Preemergence/ postemergence

Flumioxazin

Soybean and peanut

Preemergence

Fluoroglycofen

Cereal crops

Postemergence

Fluthiacet-methyl

Soybean and corn

Postemergence

Fomesafen

Soybean

Postemergence

Lactofen

Soybean

Postemergence

Oxadiargyl

Rice/sugarcane

Preemergence

Oxadiazon

Grasses and ornamentals

Preemergence/ postemergence

Oxyfluorfen

Vegetable crops

Preemergence/ postemergence

Pyraflufen-ethyl

Cereals

Postemergence

Sulfentrazone

Soybean, sugarcane, tobacco

Preemergence

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81.4.1  Absorption, Translocation, and Metabolism Most postemergence-applied Protox inhibitors are readily absorbed through the leaves (Dayan et al., 1996, 1997b;

a

Information from Senseman (2007).

700

Conductivity ( mho/cm)

600

A

B

C

500 400 300 200 100 0 0

8 16 24 32 40 48 0

8 16 24 32 40 48 0 Time (h)

8 16 24 32 40 48

Figure 81.1  Herbicidal activity of carfentrazone-ethyl and deethylated carfentrazone metabolite as measured by electrolyte leakage from leaf disks of (A) soybean (Glycine max), (B) velvetleaf (Abutilon theophrasti), and (C) ivyleaf morninglory (Ipomoea hederacea). Leaf disks were incubated in the presence of no inhibitor (•), 10 M carfentrazone-ethyl (), or the free-acid metabolite () for 20 h in darkness and then exposed to continuous light. The arrows indicate the beginning of light exposure, and the dotted lines indicate maximum conductivity from boiled samples. Data are the average of three replications  SD (from Dayan et al., 1997b).

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A

3

B

3 1

1 2

2 C

C

oxidize sulfentrazone (Dayan et al., 1996). Further studies on the biological activity of the primary metabolites of sulfentrazone showed that the initial oxidative step leading to the hydroxylation of the methyl group on the triazolinone ring did not reduce the biological activity of sulfentrazone. Further oxidation of the moiety led to the formation of less active compounds (Dayan et al., 1998).

81.4.2  Mode of Action R

R

Figure 81.2  Uptake and translocation of 14C-sulfentrazone in (A) coffee senna (Senna occidentalis) and (B) sicklepod (Cassia obtusifolia). Autoradiograms were obtained by incubating the roots of both plants in a solution containing 14C-sulfentrazone for 3 h and transferring the plants to nonlabeled solutions for another 21 h. C, cotyledons; R, root; 1, 2, and 3 refer to the first, second, and third true leaves, respectively. Notice that no radioactivity remained in the roots and that most of the radioactivity accumulated in the fully expanded leaves (1 and 2).

Ritter and Coble, 1981a). Root uptake of foliar active compounds is generally poor (Ritter and Coble, 1981b). Most of the DPE herbicides (e.g., acifluorfen, lactofen, and fluoroglycofen) are not translocated beyond the point of absorption. However, others, such as fomesafen, can be readily translocated by the xylem. Some studies demonstrate that absorption and translocation of DPE herbicides may be affected by temperature and humidity (Ritter and Coble, 1981a; Wills and McWhorter, 1981). Soil active compounds, such as sulfentrazone, are readily taken up by the roots and translocated through the xylem along the transpiration flow (Wehtje et al., 1997). Radiolabeled pulse-chase studies show that nearly all of the herbicide taken up by the roots is translocated to the foliage within 24 h (Figure 81.2) (Dayan et al., 1996, 1997c). Interestingly, related sensitive (coffee senna) and resistant (sicklepod) weeds absorbed and translocated sulfentrazone to a similar extent (see Figure 81.2). The selectivity of Protox inhibitors lies primarily in the differential rate of metabolic detoxification of the herbicide. In the case of DPE herbicides, such as acifluorfen and fluorodifen, the resistance of soybean is achieved by metabolic cleavage of the ether bridge, followed by rapid conjugation of the phenyl rings to cysteine, homoglutathione, and glucose (Figure 81.3) (Frear and Swanson, 1973; Frear et al., 1983). In the case of non-DPE structures, such as sulfentrazone, resistance is achieved via rapid oxidative degradation followed by conjugation (Dayan et al., 1996, 1997c). The nature of the conjugated metabolites was not determined in this study, but 90% of the herbicide was transformed into extremely water-soluble metabolites within 24 h from the time of application (Dayan et al., 1996, 1997c). Sensitive weeds are apparently unable to

Early investigations on the mode of action of Protox inhibitors determined that light was needed for herbicidal activity (Matsunaka, 1969); however, photosynthesis was not involved in the mechanism (Duke and Kenyon, 1986). Soon after, it was first observed that peroxidative damage was due to a process involving a photodynamic pigment (reviewed by Scalla and Matringe, 1994). Matringe and Scalla (1988) demonstrated that the chlorophyll precursor protoporphyrin IX (Proto) accumulated in diphenyl ether herbicide-treated plant tissues (Lydon and Duke, 1988; Witkowski and Halling, 1988). Proto is a photo­dynamic pigment (Cox et al., 1982), and its content in Protox inhibitor-treated plant tissues correlated well with peroxidative damage (Becerril and Duke, 1989). Matringe et al. (1989a,b) discovered that Proto accumulated in response to the inhibition of the enzyme responsible for its synthesis (Figure 81.4). This apparently contradictory situation is mirrored by the accumulation of Proto in humans with the genetic disease variegate porphyria that results from a deficiency of Protox activity (Deybach et al., 1981). The paradox of inhibition of an enzyme leading to the accumulation of its catalytic product is explained by altered compartmentalization of porphyrin intermediates (Jacobs et al., 1991; Lee et al., 1993; Lehnen et al., 1990). Inhibition of Protox induces an uncontrolled accumulation of protoporphyrinogen IX (Protogen), which leaks out of the chloroplast outer membrane into the cytoplasm, where it is converted into the highly photodynamic Proto (Figure 81.5). In the presence of light, this photosensitized Proto generates a highly reactive singlet oxygen that induces lipid peroxidation of the relatively unprotected plasma membrane (see Figure 81.5) (Devine et al., 1993; Lee et al., 1993). This phenomenon explains the light-dependent nature of the mode of action of Protox inhibitors. The catastrophic consequences of the unregulated accumulation of protoporphyrin in the cytosol and subsequent generation of reactive oxygen species also have secondary effects on other physiological processes, such as the stability of PSI and PSII reaction centers (Tripathy et al., 2007). A large number of Protox inhibitors exists, and despite their structural differences, they apparently all compete for the same site on the enzyme (Dayan et al., 1997a,b, 1998; Matringe et al., 1992), suggesting that the binding pocket is promiscuous. Figure 81.6 demonstrates the competitive

Chapter | 81  Protoporphyrinogen Oxidase-Inhibiting Herbicides

Cl

A

1739

Figure 81.3  Examples of hydrolytic and oxidative metabolic degradation of various Protox inhibitors in plants. (A) Metabolism of the diphenyl ether acifluorfen in soybean plants (Frear et al., 1983) and (B) metabolism of the triazolinone carfentrazone (Anonymous, 1995b).

COOH

F 3C

O

Acifluorfen

NO2

HOOC

+GSH

NH2 O

HOOC

Cl F 3C

O2 N

OH

NH SCH2 O

+Sugar Cl F3C

OH

O

HO O

O

COOH

OH

O

Cl F3C

HO

NH

HOOC

OH

NH2

O2 N

SCH3

COOH

OH OH CH 2COOH

O C O

F

CHF2 N

O N

B

CH3

N

Carfentrazone-ethyl

Cl

COOEt Cl

F

CHF 2 N

O N

N

Cl

F

-HCl

Esterase CHF2 N

O N

CH 3

N

CH3

CHF 2 N

O

F (O)

N

CH2OH

N

Cl

Cl

COOH

COOH Cl

Cl

COOH

F

O

(O) CHF 2 N N N

Conjugates

F

CHF 2 N

O

-CO 2

N

N

COOH

Cl

Cl

COOH

COOH Cl

binding between the natural substrate and various synthetic herbicides belonging to structurally different chemical classes. The regression of the lines intersecting with each other at the level of the y axis on a double reciprocal plot is typical of competitive binding. These graphs also enable the calculation of binding constants for each inhibitor. As would be expected, the inhibition of Protox is proportional to the ability of each compound to bind to that particular site on Protox. Early studies on DPE herbicides indicated that these molecules compete for the binding site by mimicking half of the natural substrate Protogen (Nandihalli

Cl

et al., 1992). However, no such clear resemblance to Protogen can be easily established with nonoxygenbridged Protox inhibitors (Dayan et al., 1997a–c; Reddy et al., 1998). New Protox inhibitors with even higher mimicry to Protogen have been developed. One set of compounds has structures that resemble three of the pyrrole rings of Protogen (Theodoridis et al., 1995), whereas another set, which was found to be extremely active, mimics both the hydrophobic region associated with two pyrrole rings and the hydrophilic region associated with two propionic acid side chains (Theodoridis et al., 2000).

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Glutamic acid

Chlorophyll

δ-aminolevulinic acid Chlorophyllide ALA dehydratase

Gabaculine

Protochlorophyllide Porphobilinogen PBG deaminase Uro III cosynthase

2,2’-dipyridyl Levulinate 4,6-dioxoheptanoate

Cyclase Mg-Proto methyl ester

Uroporphyrinogen III Mg-Proto

Coproporphyrinogen

Protogen

Protox Diphenyl ethers Oxadiazoles N-phenylimides Triazolinones

Proto

Heme

Figure 81.4  Simplified pathway of chlorophyll and heme biosynthesis in plants showing the enzymes known to be sensitive to chemical inhibition (boxes). ALA dehydratase, aminolevulinic acid dehydratase; PBG deaminase/Uro III cosynthase, porphobilinogen deaminase/uroporphyrinogen III cosynthase; Protox, protoporphyrinogen oxidase (redrawn from Reddy et al., 1998).

Figure 81.5  Mode of action of Protox-inhibiting herbicides. The initial step consists of the foliar absorption and translocation of the herbicide. Inhibition of Protox (localized in the outer membrane of chloroplasts) causes an unregulated accumulation of Protogen, which is oxidized (either enzymatically or chemically) to Proto. Proto is energized by light and causes formation of reactive singlet oxygen species that can lead to oxidative membrane degradation (redrawn from Dayan and Duke, 1997).

Two-dimensional quantitative structure–activity relationship (QSAR) analyses have been somewhat successful in predicting the herbicidal activities of these compounds (Nandihalli et al., 1992; Reddy et al., 1995, 1998). However, equations derived from three-dimensional QSAR have been more reliable at predicting the activity of various

structurally related groups of Protox-inhibiting herbicides, as well as differentiating active from inactive stereoisomers (Dayan et al., 1999). The publication of the crystal structures of Myxococcus xanthus Protox complexed with acifluorfen and Nicotiana tabacum mitochondrial Protox complexed with phenyl pyrazole inhibitor has advanced the current understanding of the binding of Protox inhibitors (Corradi et al., 2006; Koch et al., 2004).

81.4.3  Modes of Resistance Plants can be resistant to herbicides via physical, physiological, and/or biochemical mechanisms. Most often, natural resistance is achieved via slow uptake and translocation, rapid metabolic degradation, and/or resistance at the molecular site. The complex mode of action of Protox inhibitors provides several more unusual sites for possible herbicide resistance (Figure 81.7) (Duke et al., 1997). Reduced uptake and translocation through the shoots of foliar-applied Protox-inhibiting herbicides may account for the resistance of some species (Matsumoto et al., 1994) but plays a minor role in resistance to soil-applied Protoxinhibiting herbicides (Dayan et al., 1996, 1997b,c). On the other hand, metabolic degradation of Protox inhibitors seems to play a key role in crop resistance to these herbicides. Resistance of soybean to acifluorfen and two phenyl triazolinones is due primarily to rapid metabolic degradation of the herbicides (Dayan et al., 1996, 1997b,c; Frear et al., 1983). However, these herbicides act so rapidly that metabolic degradation does not provide a large safety margin, and some crop damage, often referred to as “bronzing,” is common. The extent of injury might be greater with those herbicides that have longer soil persistence. Rice appears to be more naturally resistant to the oxidative stress induced by Protox inhibitors relative to the targeted weeds (Matsumoto et al., 1994). That is, the plant generates Proto in response to the herbicides, but it appears to have the ability to cope with the resulting singlet oxygen and the toxic compounds (hydroxyl radical, lipid peroxides, etc.) resulting from it. We have reported that a similar mechanism may be involved in the differential sensitivity of soybean cultivars to sulfentrazone (Dayan et al., 1997c). Other species (e.g., mustards) and older tissues of some species that are sensitive in the seedling stage are apparently resistant due to enzymatic degradation of Protogen to nontoxic compounds (Jacobs et al., 1996). There are no cases of natural resistance in whole plants associated with herbicide-insensitive chloroplastic Protox. Nevertheless, herbicide-resistant Protox has been isolated from tobacco (Ichinose et al., 1995) and soybean cell cultures selected with Protox inhibitors (Pornprom et al., 1994). Overexpression of mitochondrial Protox in photomixotrophic tobacco cell lines can result in resistance to Protox inhibitors (Watanabe et al., 1998). This result suggests that if sufficient uninhibited mitochondrial Protox is available,

Chapter | 81  Protoporphyrinogen Oxidase-Inhibiting Herbicides

B

A

1741

C

2

1

0

0

1 1/nM Free

2

2

mg protein/nM bound

mg protein/nM bound

mg protein/nM bound

3

1

0

0

0.1 0.2 0.3 0.4 0.5 1/nM Free

4

2

0 0.0

0.1

0.2

0.3

0.4

1/nM Free

Figure 81.6  Competitive binding between the radiolabeled acifluorfen (DPE herbicide), Protogen (the natural substrate of Protox), and various Protox inhibitors on isolated etioplasts. The graphs illustrate the competitive nature of the binding between acifluorfen () and (A) Protogen (), (B) a triazolinone-type Protox inhibitor (), and (C) an isoxazole carboxamide-type Protox inhibitor () (redrawn, respectively, from Matringe et al., 1992; Dayan et al., 1997b; Dayan et al., 1997a).

Figure 81.7  Schematic of possible mechanisms of resistance to Protox-inhibiting herbicides. Potential sites of resistance are in boldface numbers: 1, inhibition of uptake or sequestration of the herbicide; 2, rapid metabolic degradation of the herbicide in the cytosol; 3, herbicide-resistant Protox; 4, degradation of extraplastidic Protogen and/or Proto; 5, quenching of singlet oxygen and other toxic oxygen species.

Protogen leaking from inhibited plastids can be rapidly converted to heme by the mitochondria before it can accumulate as damaging extraorganellar Proto. Whole plants have not been regenerated from any of these cell lines. The complex mechanism of action of Protox inhibitors provides several potential mechanisms for evolved resistance in weeds (see Figure 81.7). However, resistance to these herbicides has been slow to evolve in the field. This could be due, in part, to the relatively short-lived selection pressure of these fast-acting, foliar-applied herbicides. However, the development of more persistent soil-active Protox inhibitors may increase the selection pressure and consequently raises the probability of the evolution of resistance. Amaranthus tuberculatus (waterhemp, syn. A. rudis) biotypes resistant to Protox inhibitors have been identified

from Kansas, Missouri, and Illinois (Li et al., 2004; Patzoldt et al., 2005; Shoup et al., 2003). These biotypes have broad levels of cross-resistance to several Protoxinhibiting herbicide classes. Resistance levels to lactofen exceed 50-fold (relative to the wild-types) (Shoup et al., 2003; Patzoldt et al., 2006; Falk et al., 2005, 2006; Lee et al., 2008). Early studies showed that herbicide uptake, translocation, and metabolism of the herbicides were not altered between the wild-type and resistant biotypes (Shoup and Al-Khatib, 2005). However, resistance was associated with reduced accumulation of proto IX, which was accompanied by reduced membrane damage (Li et al., 2004). Resistance is an incompletely dominant trait that co-segregated with a nuclear gene designated PPX2L (Patzoldt et al., 2006). Interestingly, this gene encodes the mitochondrial PPO homolog, which is not the primary target of Protox inhibitors. However, this gene possesses an amino-terminal extension, which leads to dual targeting of the gene product toward both plastids and mitochondria (Watanabe et al., 2001). The herbicide resistance PPX2L gene was found to be missing a codon for glycine at position 210. The glycine deletion mutation may have arisen due to a slippage-like mechanism associated with short nucleotide repeats (micro­ satellites), as the deleted codon was present within such a context (Gressel and Levy, 2006; Patzoldt et al., 2006). Other cases of evolved resistance to Protox inhibitors since the initial reports of the A. tuberculatus biotypes resistant to Protox inhibitors include Euphorbia heterophylla (Trezzi et al., 2005) and Ambrosia artemisiifolia (Heap, 2008). However, the physiological, biochemical, or molecular basis for resistance to Protox inhibitors has not been elucidated in these species.

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81.4.4  Genetically Engineered Resistance The amount of herbicide applied as a percentage of the total amount of herbicide used in soybean production increased from 56% in 2000 to 86% in 2006 (USDA-NASS, 2008), which demonstrates the success of glyphosate-resistant soybeans. Approximately 90% of the soybeans and 70% of the cotton grown in the United States in 2007 were glyphosate resistant, and the percentage of maize grown with this trait is rapidly growing. Approximately 80% of the transgenic crops grown worldwide are glyphosate resistant. Because profits have been large, efforts are under way to develop other transgenic, herbicide-resistant crops. To capitalize on the potential advantages of Protox inhibitors (e.g., reduced risk classification and low rates of applications), plants resistant to Protox inhibitors have been generated by several laboratories. Transgenic tobacco plants resistant to Protox inhibitors have been generated by introducing bacterial Protox genes into plants (Choi et al., 1998; Kuk et al., 2005; Lee et al., 2000; Li and Nicholl, 2005). Tobacco and soybean cell lines resistant to photobleaching herbicides have been identified and characterized. It has been demonstrated that overexpression of both plastidic and mitochondrial Protox leads to resistance to Protox inhibitors (Lermontova and Grimm, 2000; Warabi et al., 2001; Watanabe et al., 1998). Several patents have been published on using both plant and microbial genes encoding herbicide-resistant forms of Protox to produce Protox inhibitor-resistant crops (FonnePfister et al., 2001; Horikoshi and Hirooka, 1997; Jung and Kuk, 2007; Jung et al., 2008; Lee et al., 2007; Nakajima and Nagasawa, 2005; Schiffer et al., 2001; Tanaka et al., 2007; Volrath et al., 1999, 2000a,b; Ward and Volrath, 1995, 1998; Ward et al., 2001; Yun et al., 1997). Although no transgenic crops resistant to Protox inhibitors are yet commercially available, a transgene from Arabidopsis encoding a double mutant Protox, along with a very active promoter, provides resistance to some of these herbicides in maize under field conditions (Li and Nicholl, 2005). In Arabidopsis with this gene, resistance ranged from 10 to 100 times, depending on the Protox inhibitor. It is not clear how this technology can be used to benefit a particular company because Protox inhibitor-resistant crops would be cross-resistant to most, if not all, other herbicides targeting that site. Several competing companies commercialize such agrochemicals.

81.5  Environmental impact 81.5.1  Interaction with Soil As mentioned previously, the crop selectivity of Protox inhibitors is, for the most part, limited to soybean and rice. Although no significant limitation on crop rotations has been associated with foliar-applied Protox inhibitors, there are some limitations with the more persistent soil-applied

Hayes’ Handbook of Pesticide Toxicology

Protox inhibitors (e.g., fomesafen and sulfentrazone). Nevertheless, the excellent broad-spectrum preemergence activity associated with greater soil persistence of sulfentrazone, relative to other Protox-inhibiting herbicides, makes this herbicide unique in its class for the moment. The combination of relatively high soil adsorption and rapid microbial degradation strongly limits soil leaching of most Protox-inhibiting herbicides. None of the Protox inhibitors has volatilization problems (vapor pressure lower than 107 mmHg at 25°C), and none has caused driftrelated injury to nontarget crops when properly applied. It is important to note that soil quality may affect leaching of fluoroglycofen, and metabolites of lactofen may be highly mobile in soil (Ahrens, 1994). In the case of soil active Protox inhibitors, mobility may be of some concern. The herbicide bifenox is quite mobile despite its relatively high soil sorption (Table 81.4); fortunately, its half-life is fairly short. Sulfentrazone is not as mobile, but its sorption appears to be affected by pH and soil mobility may be a problem at pH greater than its pKa (6.6) (Anonymous, 1995a; Grey et al., 1997). Typical soils of soybean fields can be as high as pH 7.5. Such pH might lead to significant levels of leaching. This problem may be compounded by the relatively long half-life of sulfentrazone (see Table 81.4).

81.5.2  Degradation in the Environment Protox inhibitors are not known to be a threat to the environment. The principal form of degradation is associated with microbial activity, although some of these herbicides (e.g., DPEs) are also susceptible to photodegradation (see Table 81.4). The half-lives of this class of herbicide vary greatly and are affected by soil quality. Half-life can be very short (i.e., less than 1 week for lactofen) but can be as long as 280 days (e.g., sulfentrazone) (see Table 81.4) (Anonymous, 1995a). In an aquatic environment, photodegradation of most Protox inhibitors is very rapid (5 days). Some Protox inhibitors, such as fluazolate (JV-485) (Prosch et al., 1997) and fomesafen (Ahrens, 1994), have low water solubility and are therefore considered a low risk to groundwater or surface water runoff and, consequently, to aquatic wildlife. Others, such as oxadiargyl (Dickmann et al., 1997), dissipate rapidly from water and do not persist in the aquatic environment. Oxadiazon, oxyfluorfen, and lactofen (Ahrens, 1994) are strongly adsorbed by the soil and therefore do not leach into groundwater, presenting a diminished toxicological risk to aquatic wildlife.

81.5.3  Ecotoxicology The majority of Protox inhibitors are not harmful to avian wildlife (Table 81.5). In general, they present a low risk to the environment (see Table 81.4) and terrestrial animals

Chapter | 81  Protoporphyrinogen Oxidase-Inhibiting Herbicides

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Table 81.4  Fate of Protox-Inhibiting Herbicides in the Environment Common name

Sorption (Koc, ml/g)

Degradation

Mobility

Volatilization

Half-life (days)

Acifluorfena

113

Photo/microbial

Negligible

Negligible

14–60

Azafenidin

n/a

Photo/microbial

Negligible

Negligible

25–40

Bifenoxa

10,000

Microbial

Significant

Negligible

7–14

2.1

Microbial

Negligible

Negligible

2-4

Carfentrazone-ethyl

n/a

Microbial

Low

Negligible

60

Fluazolateb

250

Microbial

Negligible

Negligible

40–70

n/a

n/a

n/a

n/a

n/a

Strong

Photo/hydrol.

Negligible

Negligible

1–6

Flumioxazin

n/a

Microbial

Negligible

Negligible

3-7

Fluoroglycofena

1364

Photo/microbial

Moderate

Negligible

7–21

Strong

Photolytic

Low

n/a

1-5

Fomesafen

60

Photolytic

Moderate

Negligible

100

Lactofena

10,000

Microbial

Negligible

Negligible

3–7

n/a

Microbial

Negligible

n/a

40

3200

n/a

Low

Negligible

60

Oxyfluorfen

100,000

Photo/microbial

Negligible/low

Low

30–40

Pyraflufen-ethyla

Strong

n/a

n/a

Negligible

1-7

Microbial

Moderate

Negligible

110–280

a

a

Butafenacil

a

Flufenpyr-ethyl

a

a

Flumiclorac

a

a

Fluthiacet-methyl a

a

Oxadiargyl

a

Oxadiazon

a

a

Sulfentrazone

g

160–192

a

Senseman (2007). National Registration Authority (2002).

b

(Table 81.6). Of the herbicides tested against insects, flumiclorac, fluoroglycofen, lactofen, oxyfluorfen, and fluazolate presented extremely low risk of toxicity. Fomesafen was the only inhibitor tested that exhibited moderate toxicity to bees. Aquatic wildlife is generally more susceptible to Protox inhibitors than is avian wildlife. Photodegradation times in water for most of these is very short (5 days). Some Protox inhibitors, such as fluazolate (Prosch et al., 1997) and fomesafen (Ahrens, 1994), have low water solubility and are therefore considered a low risk to groundwater or surface water runoff and, consequently, to aquatic wildlife following soil application. Others, such as oxadiargyl (Dickmann et al., 1997), dissipate rapidly from water and do not persist in the aquatic environment. When applied directly in water, oxadiazon, oxyfluorfen, and lactofen are strongly adsorbed by the soil and therefore do not leach into the groundwater, presenting a diminished toxicological risk to aquatic wildlife (Ahrens, 1994). Many Protox-inhibiting herbicides, such as carfentrazone and flumiclorac, are highly toxic to algae but are only moderately toxic to fish. Others, such as sulfentrazone and fomesafen, are essentially nontoxic to fish (Ahrens, 1994; Anonymous, 1995a). Finally, bifenox and oxyfluorfen are highly toxic to aquatic wildlife (Ahrens, 1994).

81.6  Mammalian toxicology 81.6.1  Skin and Oral Protox inhibitors have been shown to have little acute toxicity (see Table 81.6). Although the molecular target site of these herbicides was identified prior to their registration, it is now known that the target site is Protox, the last common enzyme in the biosynthesis of heme and chlorophylls. Protox-inhibiting herbicides appear to be as inhibitory to mammalian mitochondrial Protox as to that of chloroplast (Matringe et al., 1989b). These compounds increase the porphyrin blood levels in animals when administered by oral doses (see Section 81.6.3). However, the herbicides appear to be effectively metabolized and/or excreted (Adler et al., 1977; Hunt et al., 1977; Leung et al., 1991), and thus porphyrin levels return to normal within a few days. Interestingly, many of the primary mammalian metabolites formed are the same as photochemical degradation products (Hunt et al., 1977). Even under exaggerated dietary doses (100 recommended field rates), there appears to be little bioaccumulation risk in animals (Leung et al., 1991). In general, for healthy individuals, these compounds are not considered to pose any significant toxicological risk (Shaner, 2003).

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Table 81.5  Environmental Toxicology of Protox-Inhibiting Herbicides Common name a

Acifluorfen

Azafenidina

a

Bifenox

b

Butafenacil

a

Carfentrazone

Fluazolatea

a

Flufenpyr-ethyl

a

Flumiclorac

Flumioxazina

Fluoroglycofena

a

Fluthiacet-methyl

a

Fomesafen

Lactofena

Oxadiargyla

a

Oxadiazon

a

Oxyfluorfen

Avian Oral LD50 (mg/kg)

Test species

LC50 (mg/l) (96 h)

Bobwhite quail

325

Bluegill sunfish

62

Mallard duck

4187

Rainbow trout

17

Bobwhite quail

2500

Bluegill sunfish

48

Mallard duck

2500

Rainbow trout

33

Mallard duck

5000

Bluegill sunfish

0.64

Pheasant

5000

Rainbow trout

0.87

Mallard duck

674

Bluegill sunfish

2.5

Quail

674

Rainbow trout

4.8

Bobwhite quail

2250

Bluegill sunfish

2.0

Mallard duck

5620

Rainbow trout

1.6

Bobwhite quail

2130

Bluegill sunfish

0.045

Mallard duck

2130

Rainbow trout

0.045

Bobwhite quail

2250

Bluegill sunfish

2.7

—

—

Rainbow trout

3.7

Bobwhite quail

2250

Bluegill sunfish

17.4

Mallard duck

5620

Rainbow trout

1.1

Bobwhite quail

2250

Bluegill sunfish

21

Mallard duck

2250

Rainbow trout

2.3

Bobwhite quail

1075

Bluegill sunfish

1.5

Mallard duck

5000

Rainbow trout

23

Bobwhite quail

2250

Bluegill sunfish

0.14

Mallard duck

2250

Rainbow trout

0.043

Bobwhite quail

2000

Bluegill sunfish

6,000

Mallard duck

5000

Rainbow trout

680

Bobwhite quail

2510

Bluegill sunfish

560

Mallard duck

5620

Rainbow trout

0.1

Bobwhite quail

2000

Bluegill sunfish

0.37

—

—

Rainbow trout

0.37

Bobwhite quail

6000

Catfish

15.4

Mallard duck

 1000

Rainbow trout

9

Bobwhite quail

2200

Bluegill sunfish

0.2

4000

Rainbow trout

0.4

n/a

Carp

10

n/a

Rainbow trout

10

Mallard duck Pyraflufen-ethyla

a

Sulfentrazone

a

Bobwhite quail

5620

Bluegill sunfish

93.8

Mallard duck

5620

Rainbow trout

130

Senseman (2007). National Registration Authority (2002).

b

Fish

Test species

Chapter | 81  Protoporphyrinogen Oxidase-Inhibiting Herbicides

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Table 81.6  Mammalian Toxicity of Protoporphyrin Oxidase-Inhibiting Herbicides Common name

Test species

Oral LD50 (mg/kg)

Mutagenicity

Teratogenicity (mg/kg/day)

Acifluorfena

Rat Rabbit

1500 2000

Negative

n/a

Azafenidina

Rat Rabbit

5000 2000

Negative

Negative

Bifenoxa

Rat Rabbit

5000 2000

Negative

None at 200

Butafenacilb

Rat Rabbit

5000 2000

Negative

Negative

Carfentrazonea

Rat Rabbit

5000 n/a

Negative

Negative

Fluazolatea

Rat Rat

5000 5000

Negative

Negative

Flufenpyr-ethyla

Rat Rabbit

5000 n/a

Negative

Negative

Flumicloraca

Rat Rabbit

5000 n/a

Negative

1500 on rat

Flumioxazina

Rat Rabbit

5000 2000

Negative

Positive – rat Negative – rabbit

Fluoroglycofena

Rat Rabbit

1500 5000

Negative

Negative

Fomesafena

Rat Rabbit

1250–2000 n/a

Negative

Negative

Lactofena

Rat Rabbit

n/a n/a

Negative

Negative

Oxadiargyla

Rat Rabbit

5000 2000

Negative

n/a

Oxadiazona

Rat Rabbit

5000 n/a

n/a

n/a

Oxyfluorfena

Rat Rabbit

5000 5000

Positive

Toxic at 150

Pyreflufen-ethyla

Rat Rabbit

5000 n/a

n/a

n/a

Sulfentrazone

Rabbit Rat

2500 2000

Negative

Toxic at 25

a

Senseman (2007). National Registration Authority (2002).

b

The genotoxicity of sulfentrazone was investigated in fresh leukocytes from rats. Leukocytes from both mothers and pups were tested using the comet assay, and no significant genotoxic effect was induced by the herbicide. However, it was found that maternal exposure to sulfentrazone may cause some neuromuscular and behavioral deficits in the nursing pups (de Castro et al., 2007).

81.6.2  Teratogenicity and Mutagenicity All compounds have been tested under a series of mutagenicity studies, and the overwhelming weight of evidence

supports the conclusion that with the exception of oxyflu­ orfen and sulfentrazone, Protox inhibitors are not genotoxic (see Table 81.6). Teratology studies conducted on rat and rabbit have demonstrated that the majority of the compounds are not teratogenic.

81.6.3  Effects on Mammalian Porphyrin Metabolism In healthy individuals, approved Protox-inhibiting herbicides are not considered to be of significant toxicological risk due to their effects on porphyrin metabolism. To date,

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no health problems have been associated with human consumption of crops treated with these compounds (Duke and Rebeiz, 1994). Mammalian Protox is as sensitive to Protox-inhibiting herbicides as chloroplastic Protox (Birchfield and Casida, 1997; Krijt et al., 1994; Scalla and Matringe, 1994), and these compounds can cause greatly elevated levels of porphyrins in animals administered oral doses of these compounds (Krijt et al., 1994). However, these herbicides are not readily absorbed by the body during digestion and/or are rapidly degraded by metabolism (Adler et al., 1977; Hunt et al., 1977; Leung et al., 1991). In mammals, there are remarkable species differences in the levels of porphyrin accumulation resulting from exposure to Protox inhibitors, and developmental toxi­city correlates with Proto accumulation (Kawamura et al., 1996). Rats and mice seem to be particularly sensitive to Protox inhibitors. Variegate porphyria, a human disease characterized by accumulation of Proto and other porphyrins, is caused by a deficiency of Protox. Variegate porphyria-like symptoms can be generated in mice with high doses of herbicidal Protox inhibitors (Krijt et al., 1997). However, neither of two structurally divergent Protox inhibitors (oxadiazon and oxyfluorfen) affected Protox activity of the brain and liver. How this relates to human risk assessment is unknown. In general, however, relatively high doses of herbicides are required to elicit an effect, and porphyrin levels return to normal within days after withdrawal of the herbicide. Finally, Protox inhibitors have been proposed as pharmaceuticals for use in tumor phototherapy (Halling et al., 1994). Some Protox inhibitors may preferentially accumulate in tumors, resulting in sufficient differences between tumor Proto accumulation and that in adjacent tissues for exploitation in phototherapy.

81.6.4  Metabolic Degradation in Animals There are few studies that document metabolism of DPE herbicides and other Protox inhibitors in animals and wildlife. In general, the primary form of metabolite excretion is through urine and feces (Hunt et al., 1977; Leung et al., 1991). A variety of animals, including rats, rabbits, goats, sheep, cattle, and chickens, have been tested. General classes of metabolic degradation of these compounds by animals include nitro reduction, deesterification, and conjugation to glutathione, cysteine, and carbohydrates. All commercial diphenyl ether herbicides contain a pnitrophenyl substituent. In animals, this moiety is readily reduced to an amine group. Some DPE herbicides contain a carboxyester group at the meta position on the nitrophenyl ring. This ester is readily hydrolyzed to produce a very polar carboxylic acid derivative. DPEs are also easily inactivated by cleavage of the ether bridge, followed by conjugation to glutathione (Aizawa and Brown, 1999).

Hayes’ Handbook of Pesticide Toxicology

Other conjugated metabolites have been identified with glu­ curonic acid. Most of the primary metabolites are the same as those formed in plants (see Figure 81.3A). The DPE herbicides fluoroglycofen ethyl and bifenox are readily deesterified in animals. In fact, the initial steps of fluoroglycofen ethyl metabolism lead to the formation of acifluorfen (Aizawa and Brown, 1999). In animals, bifenox, following deesterification, forms 5-(2,4-dichlorophenoxy)-2-nitrobenzoic acid. The 4-hydroxyphenyl ether metabolite of nitrofen was identified as a major derivative in rabbits (Bray et al., 1953). This metabolite was excreted primarily as a glucuronide conjugate, with small amounts of free 4,4’-dihydroxydiphenyl ether also present. Another study demonstrated that rabbits further degraded these metabolites to yield their corresponding hydroxy derivatives (Matsunaka, 1976). Studies of the metabolic degradation of oxyfluorfen in animals showed that most of the metabolic products were excreted in the feces, with small amounts remaining in the urine (2% in males and 4% in females) (Adler et al., 1977). A common metabolite consisted of an amino derivative of nitro-substituted primary metabolite 4-[2-chloro-4-(triflu­ oromethyl)phenoxy]-2-ethoxybenzenamine. This amino derivative was further degraded with the consecutive conversion of the amino group to an acetamido group to yield another common metabolite (N-[4-[2-chloro-4-(trifluoro­ methyl)phenoxy]-2-hydroxy-phenyl]actamide). All non-oxygen-bridged Protox inhibitors (i.e., those structurally different from DPEs) appear to follow a similar metabolic degradation pattern. In rats and goats, most of the herbicide metabolites are found in urine, with small amounts excreted in feces and milk. In chickens, approximately 95% of the metabolites are eliminated in excreta, with small amounts (0.09%) eliminated in the eggs (Leung et al., 1991). The carboxyester group of the triazolinone herbicide carfentrazone ethyl is initially metabolized to a carboxylic acid group. Further metabolites identified in rats and lactating goats included hydroxymethylpropionic acid and cinnamic acid derivatives (Aizawa and Brown, 1999). In mammals, the propionic acid metabolites undergo further oxidation of the methyl group by the cytochrome P-450. Finally, the cinnamic acid conjugate may be further metabolized to yield a benzole acid derivative (Aizawa and Brown, 1999). Metabolism of the triazolinone herbicide sulfentrazone has been tested in rats, goats, and hens. The primary metabolite (88–95%) is 3-hydroxymethyl sulfentrazone. Other metabolites include 3-desmethyl sulfentrazone and 2,3-dihydroxymethyl sulfentrazone. Overall, triazolinone herbicides such as sulfentrazone are rapidly metabolized, with most of the compound being excreted within 3–5 days (Leung et al., 1991). Metabolic degradation of Protox inhibitors by animal systems has been reviewed by Aizawa and Brown (1999).

Chapter | 81  Protoporphyrinogen Oxidase-Inhibiting Herbicides

There are few studies that document metabolism of diphenyl ether herbicides and other Protox inhibitors in animals and wildlife. In general, the primary form of metabolite excretion is through urine and feces. A variety of animals, including rats, rabbits, goats, sheep, cattle, and chickens, have been tested. Metabolic degradation of these compounds is usually very rapid in animals, and the mechanisms of degradation are similar to those reported in plants, including nitro reduction, deesterification, and conjugation to glutathione, cysteine, and carbohydrates (El Naggar et al., 2000; Tomigahara et al., 1999). These compounds are not considered to pose any significant toxicological risk for healthy individuals, although high doses of Protox inhibitors can cause a transient accumulation of Proto in animal systems (Kawamura et al., 1996; Kritj et al., 1994, 1997).

Conclusion Protox-inhibiting herbicides may play a more important role in the future agrochemical market for several reasons. These compounds are effective at very low application rates and have generally good ecotoxicology and human toxicology profiles at recommended application rates. Many are highly compatible with the trend toward no-tillage agriculture. Furthermore, unlike with some of the other herbicide classes, weeds do not readily evolve resistance at this particular site of action, although the discovery of a Protox inhibitor-resistant Amaranthus rudis biotype suggests that resistance can and will occur if the selection pressure is sufficient. As a result, these herbicides might replace comparable products currently available for broad-spectrum weed control in soybean fields to which weeds are rapidly becoming resistant. The success of Protox inhibitors is dependent on broadening their use to include other major crops such as maize and on enhancing the resistance of crops for which these herbicides are currently being marketed. The dramatic success of transgenic glyphosateresistant crops (Duke, 2005; Duke and Powles, 2008) and the increase in glyphosate-resistant weeds associated with these crops has increased interest in transgenes for other herbicide resistances, including resistance to Protox inhibitors. Although we are aware of no evidence of any significant environmental or toxicological risks of approved Protox inhibitor herbicides, the fact that all mitochondrial Protox forms tested so far are highly sensitive might provide a clue for toxicologists (mammalian and environmental) to find overlooked effects. However, the relatively low dose rates required for herbicidal activity may be far below the dose needed to adversely affect porphyrin metabolism in animals in field situations or in humans as a result of exposures in food, water, air, or during application. This appears to the case with at least one other herbicide (glufosinate) that is equally effective on the plant and animal forms of the enzyme glutamine synthetase.

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Chapter | 81  Protoporphyrinogen Oxidase-Inhibiting Herbicides

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Matsumoto, H., Lee, J. J., and Ishizuka, K. (1994). Variation in crop response to protoporphyrinogen oxidase inhibitors. Am. Chem. Soc. Symp. Ser. 559, 120–132. Matsunaka, S. (1969). Acceptor of light energy in photoactivation of diphenyl ether herbicides. J. Agric. Food Chem. 17, 171–175. Matsunaka, S. (1976). Diphenyl ether herbicides. In “Herbicides: Chemistry, Degradation and Mode of Action” (P. C. Kearney, and D. D. Kaufman, eds.), Vol. 2, pp. 709–739. Dekker, New York. Mito, N., Sato, R., Miyakado, M., Oshio, H., and Tanaka, S. (1991). In vitro mode of action of N-phenylimide photobleaching herbicides. Pestic. Biochem. Physiol. 40, 128–135. Nakajima, H., and Nagasawa, A. (2005). Protoporphyrinogen IX oxidase variant-expressing transgenic plants resistant to weed herbicidal compounds which disrupt the porphyrin pathways of plants. U.S. 6,570,070, 88 pp. Nandihalli, U. B., and Duke, S. O. (1993). The porphyrin pathway as a herbicide target site. Am. Chem. Soc. Symp. Ser. 524, 62–78. Nandihalli, U. B., Duke, M. V., and Duke, S. O. (1992). Quantitative structure–activity relationships of protoporphyrinogen oxidaseinhibiting diphenyl ether herbicides. Pestic. Biochem. Physiol. 43, 193–211. Nandula, V. K., Reddy, K. N, Duke, S. O., and Poston, D. H. (2005). Glyphosate-resistant weeds: Current status and future outlook. Outlooks Pest Manag. 16, 183–187. National Registration Authority (2002). “Evaluation of the New Active Butafenacil”. National Registration Authority for Agricultural and Veterinary Chemicals, Kingston, Australia. Patzoldt, W. L., Tranel, P. J., and Hager, A. G. (2005). A waterhemp (Amaranthus tuberculatus) biotype with multiple resistance across three herbicide sites of action. Weed Sci. 53, 30–36. Patzoldt, W. L., Hager, A. G., McCormick, J. S., and Tranel, P. J. (2006). A codon deletion confers resistance to herbicides inhibiting protoporphyrinogen oxidase. Proc. Natl. Acad. Sci. USA 103, 12329–12334. Pornprom, T., Matsumoto, H., Usui, K., and Ishizuka, K. (1994). Characterization of oxyfluorfen tolerance in selected soybean line. Pestic. Biochem. Physiol. 50, 107–114. Prosch, S. D., Ciha, A. J., Grogna, R., Hamper, B. C., Feucht, D., and Dreist, M. (1997). JV 485: A new herbicide for pre-emergence broad spectrum weed control in winter wheat. In “Brighton Crop Protection Conference,” pp. 45-50. Reddy, K. N., Nandihalli, U. B., Lee, H. J., Duke, M. V., and Duke, S. O. (1995). Predicting activity of protoporphyrinogen oxidase inhibitors by computer-aided molecular modeling. Am. Chem. Soc. Symp. Series. 589, 221–224. Reddy, K. N., Dayan, F. E., and Duke, S. O. (1998). QSAR analysis of protoporphyrinogen oxidase inhibitors. In “Comparative QSAR” (J. Devillers, ed.), pp. 197–234. Taylor & Francis, London. Ritter, R. L., and Coble, H. D. (1981a). Influence of temperature and relative humidity on the activity of acifluorfen. Weed Sci. 29, 480–485. Ritter, R. L., and Coble, H. D. (1981b). Penetration, translocation, and metabolism of acifluorfen in soybean (Glycine max), common ragweed (Ambrosia artemissifolia), and cocklebur (Xanthium pensylvanicum). Weed Sci. 29, 474–480. Sanyal, D., and Shrestha, A. (2008). Direct effect of herbicides on plant pathogens and disease development in various cropping systems. Weed Sci. 56, 155–160. Sato, R., Nagano, E., Oshio, H., and Kamoshita, K. (1987). Diphenyletherlike physiological and biochemical actions of S-23142, a novel N-phenylimide herbicide. Pestic. Biochem. Physiol. 28, 194–200.

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Scalla, R., and Matringe, M. (1994). Inhibitors of protoporphyrinogen oxidase as herbicides: Diphenyl ethers and related photobleaching molecules. Rev. Weed Sci. 6, 103–132. Schiffer, H., Lerchl, J., Grimm, B., Lermontova, I., and Voronetskaja, V. (2001). Protoporphyrin oxidase analogs resistant to herbicidal ­ inhibitor and genes enducing them and the development of herbicide resistant plants. PCT Int. Appl. 6 WO 0136606 A2 200010525. Senseman, S. A. (ed.) (2007). “Herbicide Handbook,” 9th ed. Weed Science Society of America, Lawrence, KS. Shaner, D. L. (2003). Herbicide safety relative to common targets in plants and mammals. Pest Manag. Sci. 60, 17–24. Sherman, T. D., Duke, M. V., Clark, R. D., Sanders, E. F., Matsumoto, H., and Duke, S. O. (1991). Pyrazole phenyl ether herbicides inhibit proto­porphyrinogen oxidase. Pestic. Biochem. Physiol. 40, 236–245. Shoup, D. E., and Al-Khatib, K. (2005). Fate of acifluorfen and lactofen in common waterhemp (Amaranthus rudis) resistant to protoporphyrinogen oxidase-inhibiting herbicides. Weed Sci. 53, 284–289. Shoup, D. E., Al-Khatib, K., and Peterson, D. E. (2003). Common waterhemp (Amaranthus rudis) resistance to protoporphyrinogen oxidaseinhibiting herbicides. Weed Sci. 51, 145–150. Smith, D. A., and Hallett, S. G. (2006). Interactions between chemical herbicides and the candidate bioherbicide Microsphaeropsis amaranthi. Weed Sci. 54, 532–537. Sumida, M., Niwata, S., Fukami, H., Tanaka, T., Wakabayashi, K., and Böger, P. (1995). Synthesis of novel diphenyl ether herbicides. J. Agric. Food Chem. 43, 1929–1934. Tanaka, A., Tanaka, R., Kato, K., and Fukagawa, T. (2007). DNA and protein sequences of Synechocystis protoporphyrinogen oxidase and its use for plant acifluorfen resistance and screening of herbicides. PCT Int. Appl. WO 2007034953 A1 20070329. Theodoridis, G. (1997). Structure–activity relationships of herbicidal aryltria-zolinones. Pestic. Sci. 50, 283–290. Theodoridis, G., Baum, J. S., Hotzman, F. W., Manfredi, M. C., Maravetz, L. L., Lyga, J. W., Tymonko, J. M., Wilson, K. R., Poss, K. M., and Wyle, M. J. (1992). Synthesis and herbicidal properties of aryltriazolinones. A new class of pre- and postemergence herbicides. Am. Chem. Soc. Symp. Ser. 504, 135–146. Theodoridis, G., Bahr, J. T., Davidson, B. L., Hart, S. E., Hotzman, F. W., Baum, J. S., Hotzman, F. W., Poss, K. M., and Tutt, S. F. (1995). Alkyl 3-[2,4-disubstituted-4,5-dihydro-3-methyl-5-oxo-1 H-1,2,4triazol-1-yl)phenyl]propenoate derivatives: Synthesis and structure– activity relationships. Am. Chem. Soc. Symp. Ser. 584, 90–99. Theodoridis, G., Bahr, J. T., Hotzman, F. W., Sehgel, S., and Suarez, D. P. (2000). New generation of Protox-inhibiting herbicides. Crop. Prot. 19, 533–536. Tomigahara, Y., Onogi, M., Kaneko, H., Nakatsuka, I., and Yamane, S. (1999). Metabolism of 7-fluoro-6-(3,4,5,6-tetrahydrophthalimido)-4(2-propynyl)-2H-1,4-benzoxazin-3(4H)-one (S-53482, flumioxazin) in the rat: II. Identification of reduced metabolites. J. Agric. Food Chem. 47, 2429–2438. Trezzi, M. M., Felippi, C. L., Mattei, D., Silva, H. L., Nunes, A. L., Debastiani, C., Vidal, R. A., and Marques, A. (2005). Multiple resistance of acetolactate synthase and protoporphyrinogen oxidase inhibitors in Euphorbia heterophylla biotypes. J. Environ. Sci. Health B Pestic. Food Contam. Agric. Wastes B40, 101–109. Tripathy, B. C., Mohapatra, A., and Gupta, I. (2007). Impairment of the photosynthetic apparatus by oxidative stress induced by photosensitization reaction of protoporphyrin IX. Biochim. Biophys. Acta 1767, 860–868.

Chapter | 81  Protoporphyrinogen Oxidase-Inhibiting Herbicides

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Ward, E. R., Voltrath, S. L., Johnson, M. A., and Potter, S. L. (2001). Herbicide-tolerant protoporphyrinogen oxidase produced by DNA shuffling of plant genes. U.S. Pat. Appl. Publ. 20010016956 A1 20010823. Watanabe, N., Che, F. S., Iwano, M., Nakano, T., Takayama, S., Yoshida, S., and Isogai, A. (1998). Molecular characterization of photomixo­ trophic tobacco cells resistant to photoporphyrinogen oxidaseinhibiting herbicides. Plant Physiol. 118, 451–458. Watanabe, N., Che, F.-S., Iwano, M., Takayama, S., and Yoshida, S. (2001). Dual targeting of spinach protoporphyrinogen oxidase II to mitochondria and chloroplasts by alternative use of two in-frame initiation codons. J. Biol. Chem. 276, 20474–20481. Wehtje, G. R., Walker, R. H., Grey, T. L., and Hancock, H. G. (1997). Response of purple (Cyperus rotundus) and yellow nutsedge (Cyperus esculentus) to selective placement of sulfentrazone. Weed Sci. 45, 382–387. Wepplo, P., Birk, J. H., Lavanish, J. M., Mandredi, M., and Nielsen, D. R. (1995). 5-Aryloxybenzisoxazole esters. Am. Chem. Soc. Symp. Ser. 584, 149–160. Wills, G. D., and McWhorter, C. G. (1981). Effect of environment on the translocation and toxicity of acifluorfen to showy crotalaria (Crotalaria spectabilis). Weed Sci. 29, 397–401. Witkowski, D. A., and Halling, B. P. (1988). Accumulation of photo­ dynamic tetrapyrroles induced by aciflurogen-methyl. Plant Physiol. 87, 632–637. Yun, Y. -C., Moon, Y. -H., Choi, J. -N., Choi, K. -W., Kim, C. -H., Kim, M. -K., Guh, J. -O., and Jeon, H. -S. (1997). Herbicide-tolerant transgenic plants expressing protoporphyrinogen oxidase. Eur. Pat. Appl. 6 EP 770682 A2 19970502. Zhang, Q., Huang, M., Min, Z., and Chen, W. (2006). Research progress on herbicidally active uracil derivatives. Nongyao 45, 73–77.

Chapter 82

Chloracetanilides William F. Heydens1, Ian C. Lamb2 and Alan G. E. Wilson3 1

Monsanto Company Pioneer Hi-Bred International, Inc. 3 Pharmacia Corporation 2

82.1  Introduction

82.2.1.2  Structures

This chapter describes the toxicology of several chloracetanilide herbicides, a subclass of the acetamides [general structure R1—C(O)—N(R2R3)] that have as a common structural feature a C1H2C group as the R1 substitution. Information is provided for alachlor, acetochlor, butachlor, metolachlor, and propachlor (see Fig. 82.1 for chemical structures); all of these herbicides, with the exception of butachlor, are sold in the United States, used primarily on corn, and collectively have the largest share in this market. The herbicidal mode of action for chloracetaniiides is not totally understood. It is known that this class of herbicides inhibits the biosynthesis of lipids, alcohols, fatty acids, proteins, isoprenoids, and flavonoids. By inhibiting synthesis of various terpenoid precursors (e.g., kaurene), these herbicides appear to interfere with the production of gibberellin. Terpenes and waxes are formed via different biosynthetic pathways both using coenzyme A intermediates and substrates; interference with the synthesis of both substances may indicate a common mechanism of inhibition through actions on coenzyme A. Furthermore, it has been shown that chloracetaniiides are detoxified in plants by conjugation with glutathione. This has also led to the suggestion that these compounds cause their herbicidal effect via conjugation of acctyl coenzyme A and other ­sulfhydryl-containing enzymes, with consequent inhibition of some critical function needed for the germination or survival of seedlings.

See Fig. 82.1.

82.2.1.3  Synonyms The common name alachlor is in general use. The major trade name for alachlor products in the United States is Lasso®. The CAS registry number for alachlor is 15972-60-8.

82.2.1.4  Physical and Chemical Properties Alachlor has the empirical formula of C14H20NO2CI and a molecular weight of 269.8. It is an odorless solid at room temperature with a melting point of approximately 38°C and has a vapor pressure of 1.6  105 mm Hg at 25°C. The solubility of alachlor in water is 242 ppm at 25°C. Alachlor is also soluble in ether, acetone, benzene, chloroform, ethanol, and ethyl acetate; it is slightly soluble in heptane.

82.2.1.5  History, Formulations, and Uses Alachlor was registered and introduced in 1967 for the preplant or preemergence control of a broad spectrum of grass, sedge, and broadleaf weeds. It is used in corn, soybeans, dry beans, cotton, sorghum, sunflowers, peanuts, and other crops.

82.2  Alachlor

82.2.2  Toxicity to Laboratory Animals

82.2.1  Identity, Properties, and Uses

82.2.2.1  Irritation and Sensitization

82.2.1.1  Chemical Name Alachlor is N-methoxymethyl-2’,6’-diethyl-2-­chloroacetanilide. Hayes’ Handbook of Pesticide Toxicology Copyright © 2001 Elsevier Inc. All rights reserved

Eye and skin irritation studies conducted in rabbits showed alachlor to be nonirritating to the eye and slightly irritating to the skin. Alachlor produced skin sensitization in guinea pigs (Ahrens, 1994). 1753

Hayes’ Handbook of Pesticide Toxicology

1754

O Alachlor

N CH2CI O

O Acetochlor

N CH2CI O

O Butachlor

N CH2CI O

O Metolachlor

N

changes indicative of liver toxicity; the effect at the lowest dose was limited to a slight elevation in the liver weights of male dogs only. In a subsequent 1-year dog study, a no observable effect level (NOEL) was established at 1 mg/kg/ day based on evidence of slight anemia at 3 mg/kg/day in two animals. In the first of two chronic studies conducted in LongEvans rats, alachlor was administered in the diet at doses of 14, 42, and 126 mg/kg/day for approximately 2 years. Hepatotoxicity was evident at all dose levels. Eye examinations revealed the presence of an ocular lesion, progressive uveal degeneration syndrome (UDS). This syndrome was noted at the two highest doses tested, and may also have occurred in two rats at 14 mg/kg/day. This ocular lesion was considered to be unique to the Long-Evans rat because the response has not been observed in other strains of rats, mice, or dogs. Furthermore, the effect has not been observed in humans involved in the manufacture of alachlor (see Section 82.2.3). The second long-term feeding study in rats was conducted at dose levels of 0.5, 2.5, and 15 mg/kg/day for approximately 25 months. Liver and ocular effects were not observed at any dose level; the NOEL was 2.5 mg/kg/day based on nasal hyperplasia and submucosal gland hyperplasia at the highest dose tested. In conclusion, the lowest NOEL for all subchronic and chronic effects was determined to be 1 mg/kg/day in the 1-year dog study.

CH2CI

82.2.2.4  Pharmacokinetic Studies

O

N CH2CI

Propachior

O Figure 82.1  Structures of Acetanilides

82.2.2.2  Acute Studies Acute toxicity data have been reported by Ahrens (1994). The oral LD50 in rats ranges from 930 to 1350 mg/kg, while the dermal LD50 is reported to be 13,300 mg/kg. The 4-h inhalation LC50 in rats was shown to be greater than 5.1 mg/1, the highest concentration tested (Monsanto, 1997a).

82.2.2.3  Repeated Dose Studies Several subchronic and chronic toxicology studies of alachlor have been conducted, the results of which have rec­ ently been reported by the U.S. Environmental Protection Agency EPA (1998a) and Heydens (1998). The major studies are summarized next. Administration of alachlor to beagle dogs for 6 months at dose levels of 5, 25, 50, and 75 mg/kg/day produced

The absorption, metabolism, and excretion of alachlor has been extensively studied in rats, mice, and monkeys (EPA, 1998a; Heydens et al., 1998). Alachlor is well absorbed in rats following oral administration. The metabolism of alachlor in rats is complex due to extensive biliary excretion, intestinal microbial metabolism, and enterohepatic circulation of metabolites. In excess of 30 metabolites of alachlor have been identified in rat excreta, with approximately equal quantities appearing in urine and feces. Nearly 90% of the administered dose was eliminated in 10 days. Qualitatively, alachlor metabolism in the mouse is similar to the rat; however, there are signi­ ficant quantitative differences between the two species. In contrast, alachlor is metabolized in monkeys to a limited number of glutathione and glucuronide conjugates, which are excreted primarily via the kidney. Excretion in monkeys is more rapid than in rodents, with 90% or more of an administered dose being excreted in the urine within 48 h. The large differences in metabolic profile patterns and urinary excretion rates between rats and monkeys is thought to be due to a physiological phenomenon commonly referred to as the molecular weight threshold for biliary excretion. Being of intermediate molecular weight (i.e., 300–500 g/mol), alachlor metabolites have been shown to undergo biliary excretion and enterohepatic recirculation

Chapter | 82  Chloracetanilides

in rodents, but are not good candidates for biliary excretion in primates (Millburn, 1975; Williams, 1971). The dermal penetration of alachlor has been investigated in monkeys and shown to be relatively low. One study showed that penetration rates (i.e., percentage of applied dose that is absorbed) for the emulsifiable concentrate (EC) formulation over a 12-h period were 7.7 and 9.1% for the undiluted formulation and diluted spray solution, respectively; values of 2.7 and 5.0% were reported for the microencapsulated product (Kronenberg et al., 1988). Another study done with a spray dilution of the EC formulation reported that the penetration rate ranged from 15 to 21% after a continuous 24-h exposure (Wester et al., 1992).

82.2.2.5  Genotoxicity Studies Numerous genetic toxicology studies have been conducted that assessed a variety of in vitro and in vivo endpoints. These include studies generated for regulatory purposes using established testing guidelines conducted under good laboratory practices (GLPs) and studies published in the scientific literature, some of which have a limited amount of validation data and/or employ questionable assay conditions. The results of all these studies provide no evidence that neoplastic responses observed in the rat (discussed later) arise through a geno-toxic mode of action, and the weight of evidence indicates that alachlor does not have general genotoxic potential in mammals. The most relevant studies for evaluating genotoxic potential are well-validated, standard assays required/recommended by regulatory authorities worldwide. Alachlor has been tested in several of these test systems (Kier et al., 1996). Ames/Salmonella assays conducted on alachlor as well as urine and bile samples from alachlor-treated rats showed no mutagenic activity. Alachlor was negative in a CHO/HGPRT mammalian cell gene mutation assay when tested up to cytotoxic levels. In an in vivo cytogenetics assay, alachlor was not clastogenic to rat bone marrow cells when given orally at dose levels up to 1000 mg/kg. Alachlor was negative in rat and mouse mi-cronucleus assays conducted at doses of 600 mg/kg (ip) and 1000 mg/kg (po), respectively. In two in vivo/in vitro UDS assays, variable responses were seen at 1000 mg/kg, but alachlor was clearly negative at all other doses below that level. The 1000 mg/kg dose is near the oral LD50 and has been shown to produce severe hepatotoxicity under the conditions employed in the UDS assay. Therefore, the biological relevance of the results at this high dose is doubtful. Two specialized studies have also been conducted to assess possible interactions with nasal tissue DNA in vivo (Heydens et al., 1998). The first study was conducted to determine if alachlor bound to DNA in rat nasal tissue. Following administration of 14C-labeled alachlor at a dose of 125 mg/kg, DNA and protein were purified and harvested from nasal turbinate tissue. There was an extremely

1755

low level of net radioactivity (5 dpm/mg DNA) associated with nasal tissue DNA, the nature of the bound radioactivity was not determined, and it was difficult to eliminate the possibility that the apparent binding represented protein contamination. Nevertheless, the results do represent an upper bound on the possible level of DNA binding. The possible biological significance of this DNAassociated radioactivity was evaluated by comparing the covalent binding index (CBI) to alachlor’s oncogenic potential expressed as the TD50 (Gold et al., 1984). Using a relationship developed for hepatocarcinogens (Lutz, 1986), it was concluded that the radioactivity was much too low to be consistent with a genotoxic mode of action for the induction of nasal tumors in rats. The second study investigated the ability of alachlor to produce DNA damage following administration in the diet at a dose of 126 mg/ kg/day for 7 days. No evidence of DNA strand breakage was observed in the nasal cells. These studies support the conclusion that alachlor produces nasal tumors in rats via a mechanism that does not involve the initial induction of DNA damage. A number of studies evaluating the genotoxic potential of alachlor have been reported in the literature. General conclusions from the major studies can be summarized as follows: Alachlor has been tested for mutagenicity in 14 bacterial test systems. While one spot test and one plate incorporation test were reported positive, the other 12 tests showed a uniform and consistent pattern of negative results. Both positive and negative results have been reported for a number of in vitro studies conducted to assess chromosome effects. Some of the studies reporting positive effects were conducted using alachlor samples that were produced by inexpensive, alternative manufacturing processes that are not used by Monsanto. One of these processes was known to involve the use of the alkylating agent, chloromethyl methyl ether, which is a known mutagen and carcinogen. The test materials used in these studies (Georgian et al., 1983; Lin et al., 1987) were substantially more toxic (10- to 100-fold) to the mammalian cells tested than alachlor produced by a high-quality manufacturing process. Therefore, results from these studies are not applicable to quality-produced, name brand alachlor products and should not be included in an assessment of alachlor’s genotoxic potential. Conflicting results have been reported for two other in vitro mammalian chromosome effect studies using alachlor of high purity (Erexson et al., 1993; Meisner et al., 1992). Use of nonstandard procedures (e.g., extended treatment period) and an unusual frequency of aberrant human lymphocytes in the control group of the study reporting positive effects were undoubtedly important factors. DNA strand breakage has also been reported in vitro at concentrations eliciting cytotoxicity and which, using pharmacokinetic analyses, would be associated with lethality in vivo (Bonfanti et al., 1992). In vivo DNA strand breakage studies in rats and mice showed no evidence of

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DNA damage (Taningher et al., 1993). These studies, along with the other negative in vivo work described previously, clearly demonstrate that the in vitro DNA damage findings reported by some investigators are not reflective of the universally negative in vivo mammalian effects.

82.2.2.6  Carcinogenicity Studies The oncogenic potential of alachlor has been assessed in two bioassays conducted with the Long-Evans strain of rat and in two studies done with CD-1 mice (EPA, 1998a; Heydens, 1998). Alachlor produced significant increases in glandular stomach and thyroid follicular tumors in rats at the highest dose tested, a level exceeding the maximum tolerated dose (MTD); nasal epithelial (olfactory) tumors were also observed at lower doses. Based on these findings, the EPA had previously classified alachlor as a Group B2 carcinogen. However, the EPA’s Cancer Peer Review Committee (CPRC) reconsidered the weight of evidence for alachlor, taking into account new mechanistic information (Section 82.7) in accordance with its Proposed Guidelines for Carcinogen Risk Assessment (EPA, 1996). Alachlor was reclassified as “likely” at high doses but “not likely” at low doses to be a human carcinogen. The term “low doses” denotes anticipated human exposures resulting from pesticide use. The EPA, in the Reregistration Eligibility Decision (RED) document for alachlor, agreed that a nonlinear approach (margin of exposure, or MOE) should be used for the purpose of risk assessment (EPA, 1998a). In the first rat bioassay, alachlor was administered in the diet at dose levels of 14, 42, and 126 mg/kg/day. Surviving male rats were sacrificed after 27 months while females were sacrificed after 25 months on study. In the second study, which was conducted to follow-up on non-neoplastic effects observed in the first bioassay, rats received alachlor in their diets at dose levels equivalent to 0.5, 2.5, 15, and 126 mg/kg/day. The highest dose level of 126 mg/kg/day exceeded the maximum tolerated dose (MTD) as evidenced by excessive body weight loss (30% below controls), hepatocellular necrosis, and decreased survival. Neoplastic responses attributable to alachlor administration were observed in the nasal turbinate mucosa and glandular stomach mucosa of both sexes and in the thyroid follicular epithelium of male rats. Significant increases in stomach and thyroid tumors were restricted to the highest dose tested. One benign nasal tumor was noted at 2.5 mg/kg/day. Although single nasal tumors of this type are occasionally observed in control animals, the tumor at 2.5 mg/kg/day was considered to be treatment related by the EPA and the definitive NOEL for oncogenicity is, therefore, 0.5 mg/kg/day. Subsequent mode-of-action investigations have shown that the nasal, stomach, and thyroid tumors are produced via non-genotoxic, threshold-sensitive mechanisms. The studies supporting this conclusion are described in Section 82.7.

Hayes’ Handbook of Pesticide Toxicology

The first oncogenicity study in mice was conducted at dose levels of 26, 78, and 260 mg/kg/day for approximately 19 months. The only notable finding was the occurrence of benign lung (bronchoalveolar) tumors in high-dose females. Although the incidence was statistically different from concurrent controls, it was within the range expected for untreated animals. Due to this equivocal finding and the overall poor survival of animals in this study, a second mouse bioassay was conducted at dose levels of 20, 78, and 331 mg/kg/day for 18 months. There was no doseresponse relationship for the incidence of lung tumors and no other indication that the tumors were related to administration of the test material. Based on all the available information, it was concluded that alachlor was not oncogenic in the mouse.

82.2.2.7  Development and Reproduction Studies The reproductive and developmental toxicity database has been reported by Heydens (1998) and recently evaluated by the EPA (1998a). The EPA’s assessment included special consideration of possible effects on infants and children as required by the Food Quality Protection Act (FQPA) of 1996. This provision of the FQPA requires the use of an additional safety factor for the protection of infants and children when warranted by the severity of effects observed in toxicology studies. The EPA considered the alachlor database to be complete, and the NOELs for developmental effects were equal to or greater than those for maternal effects in both developmental toxicity studies. There fore, the EPA concluded there is no unique sensitivity from prenatal exposure. Likewise, in the reproduction study, the reproductive NOEL is greater than the systemic NOEL. Thus, no special sensitivity for infants or children was indicated. Brief details of the studies supporting these conclusions are given next. Alachlor was fed to male and female rats at doses of 3, 10, and 30 mg/kg/day throughout premating, mating, gestation, and lactation periods for three successive generations. Nephritis and an apparent decrease in ovarian weights were noted in high-dose adults. The significance of the latter finding is doubtful because there was no microscopic change in the tissue and no effects on reproductive parameters. The systemic and reproductive toxicity NOELs were 10 and 30 (or more) mg/kg/day, respectively. In a developmental toxicity study with rats, alachlor was administered by gavage at doses of 50, 150, and 400 mg/kg/day on gestation days 6–19. Maternal and fetal toxicity were noted at the highest dose tested as evidenced by maternal deaths and decreased body weight gains, a slight decrease in fetal body weight, and a slight increase in postim-plantation loss. The NOEL for developmental toxicity in the rat was 150 mg/kg/day. Oral administration of alachlor to rabbits at doses of 50, 100, and 150 mg/kg/day on gestation days 7–19 produced maternal toxicity at the highest dose tested

Chapter | 82  Chloracetanilides

but no effects on the fetus. Therefore, the NOEL for developmental toxicity in the rabbit was greater than or equal to 150 mg/kg/day.

82.2.3  Human Experience A large number of workers have had long-term occupational exposure to alachlor at the Monsanto manufacturing facility for this herbicide, which has been in continuous operation since 1968. An exposure analysis indicated that exposure of these workers exceeded that of farmers and herbicide applicators (Acquavella et al., 1994). These manufacturing workers therefore presented an opportunity to assess alachlor’s potential to produce adverse health affects in humans. Three epidemiological studies were conducted on these individuals, the focus of which was the ocular and oncogenic effects observed in the chronic rat studies. In an evaluation of ocular health, the eyes of 135 workers judged to have the highest alachlor exposure were examined for the presence of a specific eye abnormality termed pigmentary dispersion syndrome (PDS) (Ireland et al., 1994). This lesion is analogous to the initiating lesion that occurred in one specific strain of rat from the chronic alachlor studies. Eye examinations were also given to an unexposed control group. PDS was not found in any of the exposed workers, and other eye abnormalities occurred at similar rates in exposed and unexposed individuals. These results indicate that humans exposed to alachlor are not at an increased risk of developing ocular disease. Mortality and cancer incidence studies for the period 1970–1990 were originally reported in 1994 (Leet et al., 1996) and have been updated with additional data through 1993 (Acquavella et al., 1996). The mortality cohort comprised 1199 workers employed for at least 1 year between 1961 and 1993. The cancer incidence cohort was a subset of 1169 of the mortality cohort whose members lived in Iowa for some time during the period 1969–1993. Using an in-depth knowledge of the plant and process, all job titles and descriptions used in personnel records were allocated into occupational exposure categories. Each such category was assigned a qualitative exposure rank (high, medium, low, or negligible) for alachlor. A total of 1036 workers who had potential alachlor exposure met the criteria for inclusion in the mortality analysis. Mortality from all causes among workers judged to have high alachlor exposure was lower than expected (standardized mortality ratio, SMR  0.7), and mortality from cancer was similar to that for the state of Iowa (SMR  0.9). Likewise, there was no increased cancer mortality among those workers with 5 or more years of exposure and 15 or more years since first exposure, the group in which occupationally related cancers would be mostly to occur. There were no deaths from nasal, stomach, or thyroid cancers.

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There were 1025 workers who met the criteria for cancer incidence analysis, and 701 (68%) of them belonged to the high-exposure category. The cancer incidence for workers with high-exposure potential was similar to that of the Iowa population (standardized incidence ratio, SIR  1.2), especially for those exposed for 5 or more years and with at least 15 years since exposure began (SIR  1.0). There were no cases of thyroid, stomach, or nasal cancers. These results indicate that alachlor exposure had no effect on cancer incidence rates in the workers. Thus, there was no indication of increased mortality rates from cancer or any other causes among alachlor manufacturing workers with up to 25 years of follow-up. It is especially noteworthy that cancer rates were not elevated in the highest exposed alachlor workers and that there were no cases of thyroid, stomach, or nasal cancers, which were the oncogenic responses observed in the rat bioassays. The production workers were exposed to alachlor during the year at a level that exceeded that of agricultural exposure; in fact, exposure to manufacturing personnel was estimated to exceed that of pesticide applicators by a factor of 10,000 or more (Acquavella et al., 1994). Dietary exposure is even lower than that of applicators. Therefore, the absence of ocular effects, elevated mortality, and increased cancer incidence in production workers serves as an important indicator of the low potential for adverse effects among the general population, which is exposed to extremely low levels of alachlor.

82.3  Acetochlor 82.3.1  Identity, Properties, and Uses 82.3.1.1  Chemical Name Acetochlor is 2-chloro-N-ethoxymefhyl-N-(2’-ethyl6’-methylphenyl) acetamide.

82.3.1.2  Structure See Fig. 82.1.

82.3.1.3  Synonyms The common name acetochlor is in general use. The CAS registry number for acetochlor is 34256-82-1.

82.3.1.4  Physical and Chemical Properties Acetochlor has the empirical formula of C14H20NO2Cl and a molecular weight of 269.77. It is a light amber to violet-colored oily liquid at room temperature with a specific gravity of 1.110 g/ml at 30°C. Acetochlor has a boiling point of 162°C at 7 mm Hg and a vapor pressure of 3.4  108 at 25°C. Solubility in water is 233 ppm at

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25°C. Acetochlor is also soluble in organic solvents including alcohol, acetone, toluene, and carbon tetrachloride.

82.3.1.5  History, Formulations, and Uses This herbicide was registered in the United States in 1994 by the Acetochlor Registration Partnership (ARP), which is now a partnership between Monsanto Company and Dow Agro Sciences. The registration of acetochlor is owned by the ARP, but both companies compete in the marketplace with different formulations. Acetochlor is a selective herbicide that controls a broad spectrum of annual grasses, sedge, and broadleaf weeds primarily in corn. Monsanto manufactures a number of formulations under the trade name Harness®. Formulations produced by Zeneca are sold under the trade name Surpass®. Prior to formation of the ARP, two companies had separately pursued registration with independently generated toxicology databases. Therefore, there are two or more studies for each type of toxicology test performed for regulatory purposes.

82.3.2  Toxicity to Laboratory Animals 82.3.2.1  Irritation and Sensitization Acetochlor has been shown to be practically nonirritating to the eyes and skin of rabbits. A dermal sensitization study in guinea pigs was positive (Monsanto, 1997b).

82.3.2.2  Acute Studies The acute oral and dermal toxicity of acetochlor is low. The oral LD50 value is 2148 mg/kg in the rat (Monsanto, 1997b). Dermal LD50 values were shown to be greater than 2000 mg/kg in both rats and rabbits.

82.3.2.3  Repeat Dose Studies Several subchronic and chronic toxicology studies have been conducted with acetochlor. The results have been reported by the EPA (1994) and are summarized next. A 21-day dermal study in rats at dose levels ranging from 0.1 to 100 mg/kg/day resulted in mild to minimal skin irritation but no systemic effects. In a second study conducted with rabbits at doses of 100 to 1200 mg/kg/day, the NOEL for systemic toxicity was 400 mg/kg/day. The main effects in two 90-day feeding studies with rats were reductions in body weight and food consumption. The lowest observed effect levels (LOELs) in both studies were 100 mg/kg/day; the NOELs were 10 and 40 mg/kg/day. Oral administration to dogs for 12 months at 4, 12, and 40 mg/kg/day resulted in decreased food consumption, body weight loss, testicular atrophy, and an increase in relative liver weights at the highest dose tested. The NOEL

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was 12 mg/kg/day. In a second 12-month dog study, the NOEL was 2 mg/kg/day based on changes in serum chemistry values as well as renal and testicular effects in males at 10 mg/kg/day. Dietary administration of acetochlor to CD-I mice for 23 months produced excessive mortality, body weight loss, anemia, interstitial nephritis, and changes indicative of liver damage at the highest dose tested (5000 ppm); this level clearly exceeded the MTD. Increased liver, kidney, and adrenal weights were observed at lower dose levels (500 and 1500 ppm). In a second mouse study conducted at dietary levels of 10, 100, and 1000 ppm for 18 months, the NOEL for systemic toxicity was 10 ppm (1.1 mg/kg/day) in males based on renal changes and 100 ppm (13 mg/kg/ day) in females. The administration of acetochlor to rats at doses of 500–5000 ppm (26–297 mg/kg/day) for 27 months exceeded the MTD at the highest dose tested. This was evidenced by increased mortality, excessive body weight loss, hepatocellular necrosis, and other effects. A NOEL for chronic effects was not established. In two subsequent 24-month studies conducted at lower dose levels, the lowest NOEL was determined to be 7.4 mg/kg/day based on decreased body weight gain and indications of liver toxicity.

82.3.2.4  Pharmacokinetic Studies Rats were found to rapidly metabolize acetochlor to numerous polar metabolites, which were then quickly excreted in the urine and feces; more than 95% of the recovered dose was excreted within 72 h. Blood was the only tissue in which significant retention of radioactivity was observed (2–3% of the dose), and the activity was shown to be associated with hemoglobin. The major metabolite, as well as some minor degradates, were identified and shown to be a result of the mercapturic acid pathway formed by initial glutathione conjugation. The glucuronide conjugate is the major metabolite in bile, and enterohepatic recirculation is known to occur as discussed previously for alachlor. In the mouse, little or no enterohepatic recirculation occurs and the glucuronide is the major urinary metabolite. The dermal penetration of acetochlor was measured in rhesus monkeys following a 24-h exposure. The penetration of the concentrated material was 4.9%, while that of a 1:70 spray dilution was 17.3% (Wester et al., 1996).

82.3.2.5  Genotoxicity Studies Acetochlor gave uniformly negative results in a wide range of in vivo rodent genetic toxicity studies conducted in rats and mice at and below toxic levels. This inactivity is consistent with its lack of in vitro gene mutation and DNAdamaging activity. Several negative in vivo clastogenicity assays indicate that ace-tochlor’s in vitro clastogenicity is not relevant to its activity in vivo or to its ability to produce

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tumors in rodents (discussed later). This conclusion is further supported by the fact that acetochlor is not genotoxic to the olfactory nasal epithelium of rats, the primary site of its rodent carcinogenicity. Detailed discussions of the genetic toxicity data for acetochlor have been published (Ashby et al., 1996, 1997); a summary of those studies is provided next. An extensive set of studies have led to the conclusion that acetochlor is not mutagenic to Salmonella typhimurium. The experiments included observations made with strains TA98, TA100, TA1535, TA1537, and TA1538 using S9 mixes derived from rats pretreated with either aroclor or a combination of phe-nobarbitone/beta naphthoflavone. Acetochlor was inactive in an in vitro assay for unscheduled DNA synthesis (UDS) using isolated primary rat hepatocytes. In two CHO/HGPRT gene mutation assays conducted at the same concentrations, acetochlor marginally increased the mutation frequency at toxic dose levels and in the presence of S9 mix in one investigation, and produced a clearly negative response in the second assay. Acetochlor is clastogenic to human lymphocytes in vitro at cytotoxic dose levels. The nonclastogenicity of the deschloro analog of acetochlor and the clastogenicity of the desoxy (N-butyl) analog indicate that the chloroacetyl substituent on acetochlor is the clastogenic moiety. Although relatively inert, this substituent can react with sulfhydryl (—SH) groups such as that present on reduced glutathione (Ashby et al., 1996). This reactivity most likely accounts for the clastogenicity of acetochlor in isolated lymphocytes that have extremely low levels of protective glutathione. Acetochlor gave a weak positive response in the mouse lymphoma TK mutation assay when tested in the presence of S9 mix at excessively toxic levels (10% relative survival of cells). It is concluded that acetochlor is not directly mutagenic to DNA, but that it is clastogenic in vitro by virtue of the sulfur reactivity of its chloroacetyl sub-stituent. In vivo, normal levels of glutathione protect against this activity. Several in vivo assays have been conducted and provide important information regarding the relevance of the in vitro assay results. Negative results have been obtained in six separate in vivo assays assessing chromosome aberration activity in somatic and germ cells. Acetochlor was negative in an in vivo rat bone marrow cytogenetic assay, in two mouse bone marrow mi-cronucleus assays, and in two rat and a mouse dominant lethal assays. In the first rat dominant lethal assay (single oral dose), a reversible toxic effect, which was observed only at supra-MTD dose levels, led to a reduction in litter sizes at the 3-week sampling period (2000 mg/kg). This effect was not observed in either the mouse or the second rat dominant lethal assay or in multigeneration studies conducted using the dietary route up to the MTD of acetochlor. Clearly, the clastogenic activity observed in vitro was not expressed in vivo in the primary rodent cytogenetic assays.

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Acetochlor gave a negative response in the rat liver unscheduled DNA synthesis (UDS) assay when tested at dose levels up to, and including, the MTD. At a supraMTD dose level (2000 mg/kg), a weak positive UDS response was observed. However, this dose depressed hepatic glutathione levels by up to 80% and was associated with severe liver necrosis, substantial release of hepatic enzymes, and lethality among the treated animals (up to 30%). Acetochlor gave negative results in assays for DNA damage (comet assays) conducted using nasal tissue (respiratory and olfactory) derived from rats treated with the supra-MTD dose of 1750 ppm of acetochlor in the diet for either 7 days or 18 weeks. These negative data are particularly relevant, given that the primary target for acetochlor carcinogenesis in the rat is the olfactory nasal epithelium.

82.3.2.6  Carcinogenicity Studies Acetochlor has been assessed for oncogenic potential in a total of five dietary studies (three rat and two mouse). These studies have been discussed by Ashby et al., (1996) and are summarized next. Although tumors were reported at various sites, the only potentially toxicologically significant effects are nasal and thyroid tumors in the rat. In 1986, acetochlor was classified as a Group B2 carcinogen by the EPA. However, this evaluation was conducted prior to the availability of recent negative in vivo genotoxicity studies and mechanistic information. The negative in vivo genotoxicity results, along with the absence of tumors below the MTD, provide evidence that the oncogenic responses arise through non-genotoxic, threshold-sensitive mechanisms. This conclusion is further supported by the results of the mechanistic investigations (see Section 82.7). The European Union’s DGXI Labelling Committee did consider all relevant information in 1997, and concluded that acetochlor should not be classified as a carcinogen. The oncogenic potential of acetochlor in rats was assessed in one study at dietary levels of 500, 1500, and 5000 ppm (equivalent to 26, 81, and 297 mg/kg/day) and in a follow-up study at doses of 40, 200, and 1000 ppm (equivalent to 2, 10.5, and 54 mg/kg/day). The third rat bioassay was conducted at dietary levels of 8, 175, and 1750 ppm (equivalent to 0.8, 7.9, and 80 mg/kg/day). The dose of 54 mg/kg/day represented an MTD based on a 10% depression in body weight gain. A marginal increase in liver tumors was seen but only at a dose (297 mg/kg/day) that greatly exceeded the MTD. Nasal epithelial (olfactory) adenomas and small increases in thyroid follicular tumors occurred at and above the MTD. The NOEL for oncogenic effects was 26 mg/kg/day. Acetochlor’s oncogenic potential in mice was evaluated in one study conducted at dietary doses of 500, 1500, and 5000 ppm (equivalent to 85, 254, and 973 mg/kg/day) for 23 months, and in a second study at dietary doses of 10, 100, and 1000 ppm (equivalent to 1.2, 12, and 126 mg/kg/day)

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for 18 months. In the first study, the dose of 973 mg/kg/ day greatly exceeded the MTD as evidenced by decreased body weight gain (males, 70%; females, 24%), increased mortality, and organ toxicity. Under these excessively toxic conditions, the incidence of liver tumors was increased. The only consistent finding in the mouse across the two studies was a marginal increase in the incidence of lung tumors, predominantly adenomas in female mice. These tumors are common in the mouse strain (CD-I) employed in these studies and occur spontaneously at variable incidences. The incidences observed in treated animals from the acetochlor studies were within the historical control range. An apparent increase in the incidence of uterine histiocytic sarcomas was noted in all treatment groups in the first mouse study. However, there was clearly no doseresponse relationship across the 10-fold range of doses tested, and there was no increase in the second study. Based on these and other factors, the higher incidence of histiocytic sarcomas was not clearly related to acetochlor administration and is most likely the result of normal variation.

82.3.2.7  Development and Reproduction Studies Developmental toxicity has been assessed in two rat and two rabbit studies in which acetochlor was administered by gav-age (EPA, 1994). Acetochlor did not produce a teratogenic response in any of these studies. In each of the two rat studies, maternal and fetal toxicity were noted at the highest dose tested (400 and 600 mg/kg/day). The NOEL for both maternal and developmental toxicity was shown to be 150 mg/kg/day in one study and 200 mg/kg/day in the other. In a study with rabbits, acetochlor did not produce developmental toxicity at doses up to 190 mg/kg/day, the highest dose tested; the NOEL for maternal toxicity in this study was 50 mg/kg/day. The second rabbit study was conducted at levels up to 300 mg/kg/day without producing any developmental toxicity. Because the NOELs for developmental effects were equal to or greater than the NOELs for maternal effects in all three studies, it was concluded that there is no unique sensitivity from prenatal exposure. The reproductive toxicity of acetochlor has been evaluated in two 2-generation rat reproduction studies (EPA, 1994). The first study resulted in decreased body weight gain and reduced viable litter size at dietary concentrations of 1500 and 5000 ppm. The highest dose of 5000 ppm (400 mg/kg/day) exceeded the MTD. The NOEL for reproductive toxicity was 500 ppm (30.4 and 44.9 mg/kg/ day for males and females, respectively). In the second study, systemic toxicity was observed in parental animals at 160 mg/kg/day, which was the highest dose tested. Acetochlor had no effects on reproductive performance, but probably because of a slightly reduced body weight gain in offspring late in lactation at 160 mg/kg/day, the reproductive NOEL was considered to be 21 mg/kg/day.

82.4  Butachlor 82.4.1  Identity, Properties, and Uses 82.4.1.1  Chemical Name Butachlor is N-(butoxymethyl)-2-chloro-29,69-diethylacetani-lide.

82.4.1.2  Structure See Fig. 82.1.

82.4.1.3  Synonyms The common name butachlor is in general use. The CAS registry number for butachlor is 23184-66-9.

82.4.1.4  Physical and Chemical Properties Butachlor has the chemical formula C10017H26NO2Cl and a molecular weight of 311.89. It is a liquid at room temperature with a vapor pressure of 1.8  106 mm Hg at 25°C. The solubility of butachlor in water is 20 ppm at 25°C. Butachlor is miscible with alcohol, ether, acetone, and benzene.

82.4.1.5  History, Formulations, and Uses Butachlor was developed for the preemergent control of grass and broadleaf weeds in rice and barley. Butachlor is available as emulsifiable concentrate and granular formulations sold under the trade name Machete®.

82.4.2  Toxicity to Laboratory Animals 82.4.2.1  Irritation and Sensitization In studies with rabbits, butachlor was practically nonirritating to the skin and moderately irritating to the eye; dermal sensitization was observed in a study with guinea pigs (Monsanto, 1991).

82.4.2.2  Acute Studies Butachlor is slightly to practically nontoxic in standard animal tests. The oral LD50 in rats is 2000 mg/kg, and the dermal LD50 is 13,300 mg/kg (Monsanto, 1991).

82.4.2.3  Repeat Dose Studies New Zealand white rabbits were exposed to butachlor dermally for 21 days at dose levels up to 2500 mg/kg/day; the only sign of toxicity was dermal irritation, and the systemic NOEL was the highest dose tested. Feeding studies of 90 days’ duration have been performed with the Fischer 344,

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Sprague-Dawley, and Wistar strains of rats at dietary concentrations ranging from 300 to 40,000 ppm. Toxicity was manifest in one or more strains as decreased survival, body weight depression, anemia, and effects in the liver, kidney, and bladder. The lowest NOEL was observed in the Fischer rat at 300 ppm (18 mg/kg/day). A 90-day feeding study in CD-1 mice at dietary concentrations ranging from 1000 to 6000 ppm resulted in liver and kidney toxicity; the NOEL was determined to be less than 1000 ppm due to increased liver weights in male mice at the lowest dose tested. An 8-week oral capsule study was performed in beagle dogs at dose levels ranging from 10 to 100 mg/kg/day. The subchronic NOEL in dogs was 10 mg/kg/day based on indications of liver toxicity (Wilson and Takei, 1999). The chronic toxicity of butachlor has also been evaluated in dogs, mice, and rats. As in the subchronic studies, the liver, kidney, and bladder were the primary target organs in one or more species. The chronic NOELs in dogs and mice were 5 and 8 mg/kg/day, respectively. Chronic NOELs of 4 and 5 mg/kg/day were established in Fischer 344 and Sprague-Dawley rats, respectively (Wilson and Takei, 1999).

82.4.2.4  Pharmacokinetic Studies Similar to alachlor and acetochlor, investigations into the metabolism and pharmacokinetics of butachlor have revealed species differences in the way that this molecule is biotrans-formed and eliminated from the body. Butachlor metabolism in rats is complex due to extensive biliary excretion, intestinal microbial metabolism, and enterohepatic circulation of metabolites. Metabolism in rats follows three major pathways: (1) initial conjugation with glutathione followed by mercapturic acid pathway metabolism; (2) cytochrome P-450-mediated hydroxylation of the aromatic ring, its ethyl groups, and the N-butoxymethylene group; and (3) cleavage of the amide bonds via aryl amidase to form 2,6-diethylaniline, which is further oxidized to 4-amino-3,5-diethylphenol. Approximately 85% of an orally administered dose is eliminated in 48 h; 60% of the excreted material is found in feces and 40% in urine. The results of dermal penetration studies with rhesus monkeys indicate that butachlor is poorly absorbed through the skin. In studies employing a 6-h topical exposure period, only 0.02% of the dose was systemically absorbed during exposure to a granular formulation, and 5% of the dose was absorbed when an EC formulation was applied (Wilson and Takei, 1999).

82.4.2.5  Genotoxicity Studies The genotoxic potential of butachlor has been evaluated in numerous assay systems, using a variety of species, metabolic activation conditions, and endpoints. Results from an extensive battery of well-validated tests conducted under

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GLPs have shown that butachlor is not genotoxic. Butachlor produced no response in the Escherichia coli wp2 reverse mutation assay. A weak positive response was observed in the TA100 strain of Salmonella in one Ames assay; however, this finding was not reproduced in subsequent assays. When tested in cultured Chinese hamster ovary (CHO) cells, there was no mutagenic response in the HGPRT forward gene mutation assay. A CHO in vitro cytogenetics assay was also negative for clastogenicity. A bone marrow cytogenetics assay and a mouse micronucleus assay conducted at ip dose levels up to 750 and 1000 mg/kg, respectively, were both negative. Exposure of CD-1 male mice to butachlor for 7 weeks at dietary concentrations up to 5000 ppm produced significant body weight depression but no evidence of dominant lethal effects. An in vivo/in vitro DNA assay in F-344 rats at oral dose levels ranging from 50 to 1000 mg/kg produced no increase in unscheduled DNA synthesis (Wilson and Takei, 1999).

82.4.2.6  Carcinogenicity Studies Butachlor has been evaluated for oncogenic potential in two strains of rats (Sprague-Dawley and Fischer 344) and in CD-1 mice. Butachlor administration induced nasal, stomach, and thyroid tumors only in the SpragueDawley rat, but not in the Fischer 344 rat or the CD-1 mouse. In the Sprague-Dawley rat, the tumors were similar to those seen with alachlor and occurred only above toxic dose levels (i.e.,  MTD). Mechanistic studies have been conducted that support the conclusion that these tumors are not relevant in assessing the oncogenic risk to humans. The results of these investigations are presented in Section 82.7. Butachlor was administered to Sprague-Dawley rats in the diet at concentrations of 100, 1000, and 3000 ppm for 26 months. Increases in neoplastic lesions were observed in the olfactory epithelium of the nasal turbinate, glandular stomach mucosa, and thyroid follicular epithelium. The stomach tumors were observed only at the 3000-ppm level. This dose level exceeded the MTD as evidenced by reduced survival and severe body weight ­ depression (20%). Increased incidences of nasal and thyroid tumors occurred only at levels of 1000 ppm and above. No ­treatment-related tumors were noted in Fischer 344 rats fed butachlor at concentrations of 10, 100, and 1000 ppm for 24 months. Butachlor was administered in the diet to mice at concentrations of 50, 500, and 2000 ppm. This study provided no convincing evidence of oncogenic potential (Wilson and Takei, 1999).

82.4.2.7  Development and Reproduction Studies The potential of butachlor to produce developmental toxicity has been evaluated in both rats and rabbits. The ability

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of butachlor to impair normal reproduction following continuous oral exposure over two generations was evaluated in rats. The results of these studies showed that butachlor was not a terato-gen or reproductive toxin. The rat and rabbit developmental toxicity studies were conducted at dose levels ranging from 49 to 490 mg/kg/day and 50 to 250 mg/kg/day, respectively. In rats, maternal toxicity was observed at the highest dose tested, but there was no effect on the developing fetus. In the rabbit study, a slight increase in postimplantation loss and decreased fetal weights were observed at maternally toxic dose levels (150 and 250 mg/kg/day). The NOEL for maternal and fetal effects was 50 mg/kg/day. Butachlor administration at dietary concentrations of 100–3000 ppm over two successive generations did not adversely affect reproductive performance or pup survival (Wilson and Takei, 1999).

82.5  Metolachlor 82.5.1  Identity, Properties, and Uses 82.5.1.1  Chemical Name Metolachlor is 2-chloro-2’-ethyl-6’-methyl-N-(2-methoxyl-methylefhyl)acetanilide.

82.5.1.2  Structure See Fig. 82.1.

82.5.1.3  Synonyms The common name metolachlor is in general use. A code designation is CGA-24,705. The CAS registry number for metolachlor is 51218-45-2.

82.5.1.4  Physical and Chemical Properties Metolachlor has the chemical formula C10015H22NO2Cl and a molecular weight of 283.8. It is a liquid that is white to tan in color. Metolachlor has a vapor pressure of 1.3  106 and a boiling point of 100°C at 0.001 mm Hg. The solubility of metolachlor in water is 530 ppm at 20°C. Metolachlor is miscible with most organic solvents.

82.5.1.5  History, Formulations, and Uses Metolachlor was registered with the EPA in 1976. It is a selective herbicide for the control of annual grass weeds, yellow nutsedge, and some broadleaf species. Metolachlor is used in corn, peanuts, and soybeans. S-metolachlor, which contains a higher percentage of the more active of two isomers, was registered in 1997. Its formulations are soid under trade names such as Dual® Magnum®.

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82.5.2  Toxicity to Laboratory Animals 82.5.2.1  Irritation and Sensitization Metolachlor did not produce eye or skin irritation in rabbits; the material was positive in a dermal sensitization study with guinea pigs (EPA, 1995).

82.5.2.2  Acute Studies Oral (rats) and dermal (rabbits) LD50 values for metolachlor are 2780 and more than 10,000 mg/kg, respectively (EPA, 1995). Inhalation LC50 values (rats) of more than 1.75 mg/1 (EPA, 1995) and more than 4.3 mg/1 (Ahrens, 1994) have been published.

82.5.2.3  Repeat Dose Studies The results of subchronic and chronic studies have been summarized in the Reregistration Eligibility Decision (RED) document issued by the EPA (1995), and are briefly described next. A 21-day dermal study was conducted with New Zealand white rabbits at dose levels of 10, 100, and 1000 mg/kg/ day. Effects observed in high-dose animals were increased biliru-bin, liver weights (males), and kidney weights (females); the systemic NOEL was 100 mg/kg/day. A 3-month feeding study conducted in beagle dogs produced no effects at dose levels of 500 and 1000 ppm. In a 6-month feeding study with dogs, the NOEL was 300 ppm (approximately 7.5 mg/kg/day) based on decreased food consumption and body weight gain at 1000 ppm (25 mg/kg/day), the highest dose tested. Beagle dogs were fed metolachlor at dose levels of 100, 300, and 1000 ppm for 1 year. The NOEL for female dogs was 300 ppm (9.7 mg/kg/day) based on decreased body weight gain. In another chronic toxicity study, metolachlor was fed to Sprague-Dawley rats at dietary levels of 30, 300, and 3000 ppm (approximately 1.5, 15, and 150 mg/kg/day) for 2 years. At the high dose level, decreased body weight gain and increased liver weights (males) were observed. The NOEL for systemic toxicity was 300 ppm.

82.5.2.4  Pharmacokinetic Studies Data reviewed by the EPA (1995) indicate that metolachlor is readily absorbed following oral exposure and excreted in the urine and feces (approximately 50% in each component) over 3 days. Dermal absorption was assumed to be 62.8% in the human based on the results from an in vitro study with rat skin.

82.5.2.5  Genotoxicity Studies Metolachlor produced negative results in the genotoxicity assays conducted with the material (EPA, 1995). No gene

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mutations were detected in the Ames/Salmonella assay or the L5178/TK mouse lymphoma test. No chromosome aberrations were observed in a hamster micronucleus assay and a dominant lethal assay in mice. Metolachlor was negative in DNA damage/repair assays in rat liver cells and human fibroblasts. An in vivo/in vitro unscheduled DNA synthesis assay also produced no adverse effects.

for males and females, respectively) based on reduced pup weights in the F1a and F2a litters. No toxicity was observed in parental animals.

82.5.2.6  Carcinogenicity Studies

82.6.1.1  Chemical Name

The oncogenic potential of metolachlor has been assessed in two bioassays with CD-1 mice and in two studies using Sprague-Dawley rats (EPA, 1995). The EPA’s CPRC classified metolachlor as a Group C (possible human) carcinogen. This classification was based on an increased incidence of liver tumors observed in female rats, an effect that was reproduced in the second rat study. The CPRC also recommended that a margin of exposure (MOE) methodology be used for the estimation of human risk (EPA, 1995) rather than a cancer potency factor calculation. Dose levels in the 2-year mouse studies ranged from 30 to 3000 ppm in the diet. No treatment-related carcinogenic effects were noted in these studies. In a feeding study with rats, metolachlor was administered at dietary concentrations of 30, 300, and 3000 ppm. The incidence of benign liver tumors was significantly increased in high-dose females (150 mg/kg/day). This effect was reproduced in a second rat bioassay. Nasal tumors (one adenocarcinoma, one fibrosarcoma) were observed in high-dose males (vs. zero in controls); however, the EPA stated it was not clear that an obvious toxic effect was exerted on the nasal tissue (EPA, 1997).

Propachlor is 2-chloro-N-isopropylacetanilide.

82.5.2.7  Development and Reproduction Studies Two developmental toxicity studies have been conducted in Sprague-Dawley rats and another was performed using New Zealand white rabbits (EPA, 1995). The dose levels in the two rat studies ranged from 60 to 1000 mg/kg/ day. The NOELs for maternal and developmental toxicity were 300 mg/kg/day based on effects at 1000 mg/kg/day. Maternal toxicity was manifested by mortality, convulsions, and reduced body weight gain and food consumption. The effects observed in offspring were reduced mean body weight and an increase in resorptions. Rabbits were evaluated for developmental effects at doses of 36, 120, and 360 mg/kg/day. Maternal toxicity was observed at the highest dose tested, but there was no evidence of developmental toxicity at any dose level. A two-generation rat reproduction study was conducted at doses of 30, 300, and 1000 ppm in the diet (EPA, 1995). The reproductive NOEL was 300 ppm (23.5 and 26.0 mg/kg/day

82.6  Propachlor 82.6.1  Identity, Properties, and Uses

82.6.1.2  Structure See Fig. 82.1.

82.6.1.3  Synonyms The common name propachlor is in general use. The major trade name for propachlor products in the United States is Ramrod®. The CAS registry number for propachlor is 1918-16-7.

82.6.1.4  Physical and Chemical Properties Propachlor has the empirical formula of C11H14CINO and a molecular weight of 211.7. It is a tan solid with a melting point of 77°C and has a vapor pressure of 7.9  105 mm Hg at 25°C. The solubility of propachlor in water is 613 ppm at 25°C. Propachlor is soluble in most organic solvents except aliphatic hydrocarbons (Monsanto, 1995; WHO, 1993).

82.6.1.5  History, Formulations, and Uses Propachlor was introduced by Monsanto in 1965. It is a pre-emergence herbicide for annual grass and broadleaf weed control in corn and sorghum. Propachlor is available as a granular and flowable formulation; it is also available as a prepack formulation with atrazine.

82.6.2  Toxicity to Laboratory Animals 82.6.2.1  Irritation and Sensitization Eye and skin irritation studies conducted in rabbits showed propachlor to be severely irritating to the eye and slightly irritating to the skin. Propachlor produced skin sensitization in guinea pigs (EPA, 1998b).

82.6.2.2  Acute Studies Acute toxicity data have been reported by the WHO (1993). The oral LD50 in rats ranges from 550 to 1700 mg/kg, while the dermal LD50 is reported to be greater than

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20,000 mg/kg in the rabbit. The inhalation (4-h) LC50 in rats is greater than 1.2 mg/1, the maximum attainable concentration achieved in the study (EPA, 1998b).

82.6.2.3  Repeat Dose Studies Two chronic toxicity studies of propachlor have been conducted in both the rat and the mouse. No treatmentrelated effects were observed in the first study for either species (WHO, 1993), in which the highest dietary level was 500 ppm, and it was concluded that the dose level selection for these studies was inadequate. In the second study conducted in the rat (EPA, 1998b) at dietary levels up to 2500 (males)/5000 (females) ppm, the NOEL was 300 ppm (6 mg/kg/day) based on changes observed in the pyloric region of the stomach (erosion, ulceration, and hyperplasia of the mucosa and herniated mucosal glands) and centrilobular/midzonal region of the liver (hepatocellular hypertrophy and eosinophilic foci). In the second study conducted in the mouse (EPA, 1998b) at dietary levels up to 6000 ppm, the NOEL was 14.6 and 19.3 mg/ kg/day in male and female mice, respectively, based on changes observed in the glandular region of the stomach (erosion and ulceration of the mucosa and herniated mucosal glands) and centrilobular/midzonal region of the liver (hepatocellular hypertrophy and necrosis and eosinophilic foci). Administration of propachlor to beagle dogs for 1 year at dietary levels of 0, 25, 250, and 1000 ppm resulted in reductions in body weight gain and food consumption at the highest concentration (33 mg/kg/day). These effects may have resulted from poor diet palatability. The NOEL was approximately 6 mg/kg/day (EPA, 1998b). No adverse effects were observed in a 3-week dermal toxicity study of propachlor conducted in rats; the no observed adverse effect level (NOAEL) was 500 mg/kg/day (Rush, 1998).

82.6.2.4  Pharmacokinetic Studies The absorption, distribution, metabolism, and excretion of propachlor has been studied extensively in the rat (WHO, 1993). Propachlor is well absorbed following oral administration; following a single oral dose, approximately 70% is recovered in urine 48–56 hours after administration. The biotrans-formation of propachlor is complex due to extensive biliary excretion, intestinal microbial metabolism, and enterohepatic recirculation of metabolites. Propachlor is initially metabolized via the mercapturic acid pathway; the molecule is conjugated to glutathione and excreted in the bile along with the catabolites cysteinyl-glycine, cysteine, and N-acetylcysteine-mercapturic acid. The biliary mercapturic acid metabolites undergo de-conjugation via intestinal/microbial carbon-sulfur (C-S) lyase activity and can be reabsorbed. The reabsorbed metabolites are subsequently glucuronidated and eliminated in the urine or

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bile. Glucuronides eliminated in the bile can subsequently undergo further enterohepatic recirculation. Elimination of propachlor is rapid; more than 90% of a single dose is excreted within 48 h, primarily in the urine. Dermal penetration studies of propachlor have not been conducted. However, such data are available for one liquid formulated product; the dermal penetration rates were 20% and 51% for undiluted and diluted spray formulations, respectively (van de Sandt, 2000).

82.6.2.5  Genotoxicity Studies Negative results were observed in in vitro prokaryote (Ames/Salmonella) and mammalian gene mutation (CHO/ HGPRT) assays of propachlor (WHO, 1993). Propachlor had no effect on DNA repair in rat hepatocytes following the conduct of in vitro and in vivo/in vitro UDS assays (EPA, 1998b). A possible weak clastogenic response was observed in an in vitro mammalian (CHO) assay in the presence of metabolic activation; no evidence of clastogenic activity was observed in the absence of a metabolic activation system (EPA, 1998b). In vivo, there was no evidence of a clastogenic response in a rat bone marrow cytogenetics assay (EPA, 1998b). Propachlor had no effects on germ cells in a dominant lethal study (EPA, 1998b) conducted in rats at dietary concentrations up to 2500 ppm (111.8 mg/kg/day). Overall, the weight of evidence indicates that propachlor is not genotoxic or clastogenic in mammals.

82.6.2.6  Carcinogenicity Studies Two carcinogenicity studies of propachlor have been conducted in both the rat and the mouse (WHO, 1993). No treatment-related carcinogenic effects were observed in the first study for either species in which the highest dietary level was 500 ppm. In the second study conducted in the rat (EPA, 1998b) at dietary levels up to 2500 (males)/5000 (females) ppm, the only evidence of a possible oncogenic effect was a single carcinoma observed in the pyloric (nonglandular) region of the stomach in one male only, at the highest dietary level (2500 ppm). No carcinogenic effects were seen in the female stomach at twice this dietary concentration (5000 ppm). Erosion and proliferation of the pyloric mucosa were observed microscopically at this dietary concentration (2500/5000 ppm); these findings were dose dependent, showed a clear threshold, and are consistent with a non-genotoxic mode of action. In the second study conducted in the mouse (EPA, 1998b) at dietary concentrations up to 6000 ppm, a statistically significant increase in hepatic adenomas was observed in males at the highest dietary level only. Administration of propachlor to male CD-1 mice in their diet for a period of 3 months produced a statistically significant increase in hepatic cell proliferation at 6000 ppm (Hotz and Wilson, 1998); the

Chapter | 82  Chloracetanilides

increased cell proliferation was shown to be dose dependent with a clear threshold at 1000 ppm (NOEL). The increase in hepatic cell proliferation is believed to represent a regenerative response to the underlying severe nonneoplastic changes observed in this organ (see Section 82.6.2.3). These data provide support for a non-genotoxic, threshold-sensitive mechanism (cell proliferation) being responsible for the increase in the predominantly benign hepatic tumors observed in male mice at a dietary concentration of 6000 ppm.

82.6.2.7  Development and Reproduction Studies Two reproduction studies have been conducted in the rat with propachlor; in the first study (EPA, 1998b), parental toxicity was observed at the high dose level only (30 mg/ kg/day), by reductions in food consumption and body weight and microscopic changes to the liver (eosinophilia and hypertrophy). No treatment-related reproductive effects were observed in the study and the NOELs for systemic and reproductive toxicity were 3 and 30 (or more) mg/kg/day, respectively. In the second reproduction study (EPA, 1998b), the parental MTD was clearly exceeded at the highest dietary concentration of 2500 (males)/ 5000 ppm (females); offspring survival and weights were also adversely affected and this dietary level was discontinued after the first litter was weaned. Parental toxicity was observed at a dietary level of 1000 ppm by reductions in weight gain and a microscopic change in the liver (hepatocyte hypertrophy). Slight effects on offspring weight were observed late in the lactation period at a dietary concentration of 1000 ppm. The NOEL for systemic toxicity to parents and offspring was 100 ppm (7 mg/kg/day), and the NOEL for reproductive effects was 1000 ppm (75 mg/kg/ day). In a developmental toxicity study in the rabbit (EPA, 1998b), propachlor was administered by gavage at doses of 5.8, 58.3, and 116.7 mg/kg/day on gestation days 7–19. The maternal MTD was exceeded at the highest dose, as indicated by mortality, clinical signs, body weight loss, and reduced food consumption. Slight effects on postimplantation loss, the number of viable fetuses, and fetal weight were observed at the highest dose. None of the effects was statistically significant, and all values were within the relevant laboratory historical control ranges; however, they were considered to be possible effects of treatment. The NOEL for maternal and developmental toxicity was 58.3 mg/kg/day. In a developmental toxicity study in the rat (EPA, 1998b), propachlor was administered by gavage at doses of 20, 60, and 200 mg/kg/day on gestation days 6–19. The NOEL was 200 mg/kg/day. Propachlor had no effect on germ cells in male rats at a dietary concentration up to 2500 ppm (111.8 mg/kg/day). Overall, the weight of evidence indicates that propachlor does not produce developmental or reproductive effects.

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82.6.3  Human Experience Effects reported in humans occupationally exposed to propachlor have been limited to local skin changes (Von Schubert, 1979). Positive responses have been observed following controlled skin patch testing (Iden and Schroeter, 1977). This is consistent with observations made in experimental animals.

82.7  Mode-of-action evaluations: oncogenicity As discussed previously, chronic administration of chloracetanide herbicides has resulted in the production of nasal, stomach, liver, and thyroid tumors in rats. Most tumors occurred at excessively toxic dose levels, at or above the MTD. Such oncogenic responses are of questionable significance because of the extreme doses required to produce them. A weight-of-evidence analysis of mutagenicity results indicates that chloracetanides have no significant genotoxic potential in mammalian systems. This suggests that the oncogenic responses in rats arise through nongenotoxic, threshold-sensitive mechanisms. To better understand the relevance of the rat nasal, thyroid, and stomach tumors to humans, extensive mechanistic investigations were undertaken with alachlor, acetochlor, and butachlor.

82.7.1  Rat Nasal Tumors 82.7.1.1  Alachlor Two metabolic pathways have been proposed to explain the development of chloracetanilide-induced nasal tumors in rats. The first scheme involves the generation of formaldehyde, while the second pathway leads to quinone mine metabolite formation. The formaldehyde theory has not been supported by the available data and is only briefly mentioned here. This is followed by a discussion of the quinone mine pathway, which is widely accepted as being involved in nasal tumor induction. Alachlor and other chloracetanllfde herbicides undergo O-de-methylation with the release of formaldehyde (Brown et al., 1988; Jacobsen et al., 1991). This led to the suggestion that formaldehyde may be involved in the nasal carcinogenicity of alachlor and other chtoracetanilide herbicides. However, although formaldehyde is known to produce rat nasal tumors upon inhalation exposure, the nature and distribution of these tumors are quite different from those seen with the chloracetanilide herbicides. For example, formaldehyde-induced nasal lesions are essentially confined to the anterior nose, in regions lined by transitional or respiratory epithelium (Morgan et al., 1986), whereas chloracetanilide-induced lesions are essentially

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confined to the posterior region, lined by olfactory epithelium (Morgan et al., 1997). In addition, formaldehyde nasal tumors are characterized by marked irritancy and involve carcinomas of the squamous epithelium, the squamous cell metaplasia being a consequence of the severe damage produced in the nasal epithelium due to the irritant properties of the molecule. In contrast, there is no evidence of irritancy with the chloracetanilides, and the tumors are predominately benign adenomas of the olfactory epithelium. In addition, as mentioned previously, primary O-deatkylation of alachlor occurs in the liver and there is little evidence of unchanged alachlor reaching the nasal turbinate area. Thus, any formaldehyde released during the metabolism of alachlor in the rat liver would be expected to rapidly undergo further metabolism and not be available in the systemic circulation to reach the offactory area. Therefore, the available evidence does not support formaldehyde as the metabolite responsible for the nasal carcinogenicity observed with chloracetanilide herbicides. Investigations to understand the mechanism by which chloracetanilides produce nasal tumors in rats were initially undertaken with alachlor. Early whole-body autoradiography (WBA) work showed that alachlor-derived radioactivity specifically localized in the nasal mucosa of rats but not mice or monkeys. In another WBA study, the tertiary amide methylsulfide metabolite of alachlor, a metabolite arising only through enterohepatic circulation, was orally administered to rats. Specific localization in nasal mucosa was again observed. The intensity of the labeling was greater than that observed when parent alachlor was administered. These findings showed the importance of metabolic processes in the localization of alachlor metabolites in rat nasal tissue and provided evidence implicating metabolism in the production of nasal tumors. Significant species differences in metabolism were demonstrated that provided a mechanistic basis for the rat-specific production of nasal tumors. In rats, alachlor is initially metabolized primarily in the liver via the P-450 pathway and by glutathione conjugation (Feng and Patanella, 1988, 1989). The glutathione conjugates and their metabolites undergo enterohepatic circulation with further metabolism in liver and nasal tissue to form the putative carcinogenic metabolite, a diethyl quinoneimine (DEIQ). Higher rates of intestinal microbial metabolism, enterohepatic circulation, and target tissue metabolism result in a greater conversion of alachlor to DEIQ in rat nasal mucosa as compared to other species (Feng et al., 1990). Quinone imines such as DEIQ are electrophilic, deplete glutathione (GSH), and can exert toxicity by binding to cellular proteins. It has been shown that a DEIQ protein adduct is the major alachlor-derived protein adduct in rat nasal mucosa (Wilson et al., 1995a); however, no evidence of this adduct was observed in the nasal tissue of mice or monkeys (Heydens et al., 1998). The binding of DEIQ to nasal protein is thought to disturb cell structure and function,

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which leads to cytotoxicity, prolonged regenerative cell proliferation, and the eventual development of nasal tumors. Increased cell proliferation was shown in rats but not mice, and a clear threshold was demonstrated; the proliferation was sustained during treatment and reversible after dosing was terminated (Heydens et al., 1998). These findings indicate that increased cell turnover is a prerequisite for tumor development and that its induction is threshold sensitive. Critical differences in metabolic capability result in much higher formation of DEIQ in the nasal mucosa of rats than other species. For example, the ability of rat nasal tissue to convert the secondary sulfide metabolite of alachlor to 2,6-DEA-phenol, the proximate metabolite of DEIQ, is more than 30 times greater than that of monkeys (Li et al., 1992) and 751 times higher than that of human nasal tissue (Wilson et al., 1995b). Further, species differences are even greater when the relative enzymatic rate of alachlor conjugation to GSH (the initial metabolic step) is included in the overall rate of DEA-phenol formation from alachlor; the overall ability of rats to convert alachlor to DEA-phenol is 3000- and 22,000-fold greater than that of humans when the initial GSH conjugation occurs in the liver or nose, respectively (Wilson et al., 1995b). These data indicate that the potential for the formation of the reactive DEIQ metabolite in human nasal tissue is negligible. The results further indicate that the rat is not an appropriate model for assessing alachlor’s oncogenic risk to humans.

82.7.1.2  Butachlor The metabolism of butachlor closely parallels that of alachlor. The only difference between these two molecules is the length of the N-alkoxymethyl side chain. An important, common step in the metabolism of alachlor and butachlor is P-450-mediated N-dealkylation of this side chain. Once this occurs, the product of the two parent molecules is identical, and the subsequent metabolism to DEIQ and toxicologic response would be the same. Indeed, it has been shown that oral administration of butachlor results in the same rat-specific nasal localization and induction of cell proliferation observed with alachlor. Similar differences in nasal metabolism across species as observed with alachlor are also apparent for butachlor; with the potential for DEIQ formation substantially higher in the rat.

82.7.1.3  Acetochlor The metabolism of acetochlor has also been extensively studied. These studies have confirmed that the overall metabolism of acetochlor shares critical commonality with alachlor and butachlor. In rats, metabolism in the liver and gastrointestinal tract produces sulfur-containing metabolites that are delivered to the nose and undergo further metabolism. WBA studies corresponding to those described previously for alachlor have been conducted

Chapter | 82  Chloracetanilides

for acetochlor. Acetochlor-derived radioactivity localized in the nasal turbinates of rats but not in mice. Analysis of protein adducts found in nasal tissue from rats treated with acetochlor or its methylsulfide metabolite confirmed that the major adducts involved EMIQ, the acetochlor quinone imine metabolite analogous to DEIQ from alachlor. As with alachlor, these adducts were not found in the nasal tissues of mice or monkeys. Similarly, when acetochlor was fed to rats at doses that produced nasal tumors in the chronic studies, prolonged cell proliferation was observed. Acetochlor administration did not induce cell proliferation at the same dose level in mice. Acetochlor administration to rats at 200 ppm, a nononcogenic dose level, did not cause increased cell proliferation. In vitro metabolism studies similar to those done with alachlor have measured the enzymatic activities involved in the conversion of acetochlor and key metabolites to EMIQ. The data showed that the potential for EMIQ formation is significantly lower in the mouse and monkey than the rat. Overall, the relative rates for conversion of acetochlor to quinone imines in rat, mouse, and monkey were very similar to those seen with alachlor, suggesting that the differences between rat and human nasal tissue for acetochlor would also be comparable to that seen with alachlor.

82.7.2  Rat Stomach Tumors Butachlor and alachlor are close structural analogs that produce the same stomach tumors in rats. An extensive mechanistic research program was undertaken to understand the mechanism by which these tumors are induced. A stomach tumor initiation–promotion study demonstrated that butachlor was not active as an initiator, but it did promote the formation of tumors after treatment with N-methyl-N-nitro-N-nitrosoguanidine (MNNG), a known initiator (Branch et al., 1995). A subsequent tumor promotion study with alachlor showed that it, too, produced stomach tumors by the same promotional activity. The results of these studies provided direct experimental evidence indicating that the stomach tumors are produced by a non-genotoxic mode of action. Mechanistic studies with butachlor have shown that chloracetanilide-induced gastric neoplasia involves toxicity (atrophy) to the fundic mucosa as an initial event following high dose exposure (Hard et al., 1995; Thake et al., 1995). This atrophy then results in compensatory cell proliferation in the fundic mucosa. The accompanying profound loss of parietal cells leads to an extensive gastric hypochlorhydria and a subsequent increase in pH of the gastric contents. This increase causes excessive gastrin production, resulting in a substantial elevation of serum gastrin levels. The tropic effect of long-term stimulation of enterochromaffin-like (ECL) cells and fundic stem cells by gastrin

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further drives a sustained regenerative cell proliferation response that ultimately results in the induction of the gastric neoplasms observed in the chronic studies. Additional work demonstrated that high-dose alachlor exposure also induces the same mucosal atrophy, hypochlorhydria, and hyper-gastrinemia that characterize the unique oncogenic mechanism demonstrated with butachlor. Mucosal atrophy did not occur at a lower, nononcogenic dose of alachlor. Prolonged exposure to toxic doses of alachlor and butachlor were required to produce stomach tumors in rats. Such exposure would not occur in humans; results from a study with rhesus monkeys showed that mucosal atrophy, the initial pre-neoplastic event, did not occur at doses ranging from 100 to 400 mg/kg/day. These doses are comparable to, and higher than, the oncogenic dose (150 mg/kg/day) in rats. Based on all the data, it was concluded that these chloracetanilide-induced stomach tumors are not relevant to humans (Heydens et al., 1998).

82.7.3  Rat Thyroid Tumors Various studies have shown that the prolonged alteration of thyroid homeostasis can lead to the development of thyroid follicular tumors in rats (Hill et al., 1989; Thomas and Williams, 1991; Zbinden, 1989). The oncogenic response is mediated via increased levels of circulating thyroid stimulating hormone (TSH), which result in hyperplasia and, ultimately, neoplasia. One mechanism causing such a thyroid imbalance involves the induction of liver enzymes (McClain, 1989). This induction increases the rate of thyroid hormone excretion and is responsible for the compensatory elevation in TSH observed. Separate studies were conducted to determine if alachlor, acetochlor, and butachlor produce thyroid tumors by this mechanism (Ashby et al., 1996; Wilson and Takei, 1999; Wilson et al., 1996). The results of these studies showed that the administration of each chloracetanilide at the dose level producing thyroid neoplasia in the chronic studies caused significant increases in liver weight and T4-UDPGT (thyroxine-uridine diphosphate glucuronosyl transferase) activity, serum TSH levels, and thyroid weight at several time points. These changes were observed as early as 7–14 days after dosing began and continued throughout 2 or more months of dosing. Reversibility studies with alachlor and butachlor showed that serum TSH and hepatic UDPGT activity returned to normal after dosing was discontinued. Rats are especially sensitive to altered thyroid function by this mechanism because of their susceptibility to liver enzyme induction and lack of thyroid binding globulin in plasma, making thyroid hormone more susceptible to metabolic activity. Furthermore, because this mechanism is believed to be a threshold-sensitive phenomenon, it is not expected to be relevant for humans under actual exposure scenarios with chlo-racetanilide herbicides.

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82.8  Common mechanism of toxicity The Food Quality Protection Act of 1996 requires the EPA to perform a combined risk assessment for chemicals that produce adverse effects by a common mechanism of toxicity. The extensive database of mechanistic information, developed (see Section 82.7) to support (1) a nongenotoxic threshold mechanism of action and (2) lack of relevance to humans for the nasal turbinate, stomach, and/or thyroid oncogenic effects produced in rats for alachlor, acetochlor, and butachlor, provides the basis for considering whether these chemicals can be grouped based on a common mechanism of toxicity. The data support grouping alachlor, acetochlor, and butachlor with respect to a common mechanism of toxicity for nasal turbinate and thyroid tumors, and alachlor and butachlor for stomach tumors.

References Acquavella, J. F., Ireland, B., Leet, T., Anne, M., Farrell, T., and Martens, M. (1994). Epidemiological studies of morbidity and mortality among alachlor manufacturing workers. In “Proceedings of the XII Joint CIGR, IAAMRH, IUFRO International Symposium: Health, Safety and Ergonomic Aspects in Use of Chemicals in Agriculture and Forestry,” pp. 184–194. Acquavella, J. R., Riordan, S. G., Anne, M., Lynch, C. R., Collins, J. J., Ireland, B. K., and Heydens, W. F. (1996). Evaluation of mortality and cancer incidence among alachlor manufacturing workers. Environ. Health Perspect. 104, 728–733. Ahrens, W. H. (1994). “Herbicide Handbook of the Weed Science Society of America,” 7th ed. Weed Sci. Soc. Am, Champaign, IL. Ashby, J., Kier, L., Wilson, A. G. E., Green, T., Lefevre, P. A., Tinwell, H., Willis, G. A., Heydens, W. F., and Clapp, M. J. L. (1996). Evaluation of the potential carcinogenicity and genetic toxicity to humans of the herbicide acetochlor. Hum. Exp. Toxicol. 15, 702–735. Ashby, J., Tinwell, H., Lefevre, P. A., Williams, J., Kier, L., Adler, I.-D., and Clapp, M. J. L. (1997). Evaluation of the mutagenicity of acetochlor to male rat germ cells. Mutat. Res. 393, 263–281. Bonfanti, M., Taverna, P., Chiappetta, L., Villa, P., D’lncalci, M., Bagnati, R., and Fanelli, R. (1992). DNA damage induced by alachlor after in vitro activation by rat hepatocytes. Toxicology 72, 207–219. Branch, D. K., Shibata, M., Thake, D. C., “Wilson, A. G. E. (1995).” Gastric Tumor Initiation/Promotion Study of Butachlor in SpragueDawley Rats. and Presented at Annual Conference of International Federation of Societies of Toxicologic Pathologists, Tours, France. Brown, M. A., Kimmel, E. C., and Casida, J. E. (1988). DNA adduct formation by alachlor metabolites. Life Sci. 43, 2087–2094. Erexson, G. L., Bryant, M. R., Doer, C. L., Kwanyuen, P., and Kligerman, A. D. (1993). Cytogenetic analyses of human peripheral blood lymphocytes exposed to alachlor in vitro. Environ. Mot Mutagen. 21, 19. Feng, P. C. C., and Patanella, J. E. (1988). Identification of mercapturic acid pathway metabolites of alachlor formed by liver and kidney homogenates of rats, mice, and monkeys. Pestic. Biochem. Physiol. 31, 84–90.

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Feng, P. C. C., and Patanella, J. E. (1989). In vitro oxidation of alachlor by liver microsomal enzymes from rats, mice, and monkeys. Pestic. Biochem. Physiol. 33, 16–25. Feng, P., Wilson, A., McClanahan, R., Patanella, J., and Wratten, S. (1990). Metabolism of alachlor by rat and mouse liver and nasal turbinate tissues. Drug Metab. Dispos. 18, 373–377. Georgian, L., Moraru, I., Draghicescu, T., Dinu, I., and Ghizelea, G. (1983). Cytogenetic effects of alachlor and mancozeb. Mutat. Res. 116, 341–348. Gold, L. S., Sawyer, C. B., Magaw, R., Backman, G. M., deVediana, M., Levin-son, R., Hooper, N. K., Havender, W. R., Bernstein, L., Peto, R., Pike, M. C., and Ames, B. N. (1984). A carcinogenic potency data base of the standardized results from animal bioassays. Environ. Health Perspect. 58, 9–319. Hard, G. C., Iatropoulos, M. J., Thake, D. C., Wheeler, D., Tatematsu, M., Hagiwara, A., Williams, G. M., and Wilson, A. G. E. (1995). Identity and pathogenesis of stomach tumors in Sprague-Dawley rats associated with the dietary administration of butachlor. Exp. Toxicol. Pathol. 47, 95–105. Heydens, W. R. (1998). Summary of toxicology studies with alachlor. J. Pestic. Sci. 24, 75–82. Heydens, W. R., Wilson, A. G. E., Kier, L. D., Lau, H., Thake, D. C., and Martens, M. A. (1998). An evaluation of the carcinogenic potential of the herbicide alachlor to humans. Hum. Exp. Toxicol. 18, 363–391. Hill, R. N., Erdreich, L. S., Paynter, O. E., Roberts, P. A., Rosenthal, S. L., and Wilkinson, C. R. (1989). Thyroid follicular cell carcinogenesis. Fundam. Appl. Toxicol. 12, 629–698. Hotz, K. J., Wilson, A. G. E. (1998). “Effect of Propachlor on Cell Proliferation in the Liver of Male Mice.” Monsanto Unpublished Report. Iden, D. L., and Schroeter, A. L. (1977). Allergic contact dermatitis to herbicides. Arch. Dermatol. 113, 983. Ireland, B., Acquavella, J., Farrell, T., Anne, M., and Fuhremann, T. (1994). Evaluation of ocular health among alachlor manufacturing workers. J. Occup. Med. 36, 738–742. Jacobsen, N. E., Sanders, M., Toia, R. P., and Casida, J. E. (1991). Alachlor and its analogues as metabolic progenitors of formaldehyde: Fate of N-methoxymethyl and other N-alkoxylalkyl substltuents. J. Agric. Food Chem. 39, 1342–1350. Kier, L. D., Heydens, W. R., Lau, H., Thake, D. C., and Wilson, A. G. E. (1996). Genotoxicity studies of alachlor. Toxicologist 30, 231. Kronenberg, J. M., Fuhremann, T. W., and Johnson, D. E. (1988). Percutaneous absorption and excretion of alachlor in rhesus monkeys. Fundam. Appl. Toxicol. 10, 664–671. Leet, X., Acquavella, J., Lynch, C., Anne, M., Weiss, N., Vaughan, T., and Checkoway, H. (1996). Cancer incidence among alachlor manufacturing workers. Am. J. Ind. Med. 30, 300–306. Li, A. A., Asbury, K. J., Hopkins, W. E., Feng, P. C., and Wilson, A. G. E. (1992). Metabolism of alachlor by rat and monkey liver and nasal turbinate tissue. Drug Metab. Dispos. 20, 616–618. Lin, M. R., Wu, C. L., and Wang, T. C. (1987). Pesticide clastogenicity in Chinese hamster ovary cells. Mutat. Res. 88, 241–250. Lutz, W. K. (1986). Quantitative evaluation of DNA binding data for risk estimation and for classification of direct and indirect carcinogens. J. Cancer Res. Clin. Oncol. 112, 85–91. McClain, R. (1989). The significance of hepatic microsomal enzyme induction and altered thyroid function in rats: Implications for thyroid gland neoplasia. Toxicol. Pathol. 17, 294–306.

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Meisner, L. P., Belluck, D. A., and Roloff, B. D. (1992). Cytogenetic effects of alachlor and/or atrazine in vivo and in vitro. Environ. Mol. Mutagen. 19, 77–82. Millburn, P. (1975). Excretion of xenobiotic compounds in bile. In “The Hepatobiology System: Fundamental and Pathological Mechanisms” (W. Taylor ed.), p. 109. Plenum, New York. Monsanto Company (1991). “Material Safety Data Sheet: Butachlor Technical.” Monsanto Company, St. Louis, MO. Monsanto Company (1995). “Material Safety Data Sheet: Propachlor Technical.” Monsanto Company, St. Louis, MO. Monsanto Company (1997a). “Material Safety Data Sheet: Alachlor Technical.” Monsanto Company, St. Louis, MO. Monsanto Company (1997b). “Material Safety Data Sheet: Acetochlor Technical.” Monsanto Company, St. Louis, MO. Morgan, K. T., Gross, E. A., Joyner, D. R., Ishmael, J., and Thake, D. (1997). Proliferative nasal lesions induced in rats by alachlor, acetochlor, and butachlor originate in specific regions of the olfactory mucosa. Toxicologist 36, 12. Morgan, K. T., Jiang, X. Z., Starr, T. B., and Kerns, W. D. (1986). More precise localization of nasal tumors associated with chronic exposure of F-344 rats to formaldehyde gas. Toxicol. Appl. Pharmacol. 82, 264–271. Rush, R. E. (1998). “Propachlor: A 21-Day Dermal Toxicity Study in Rats.” Monsanto Unpublished Report. Taningher, M., Terranove, M. P., Airoldi, L., Chiappetta, L., and Parodi, S. (1993). Lack of alachlor induced DNA damage as assayed in rodent liver by the alkaline elution test. Toxicology 85, 117–122. Thake, D. C., Iatropoulos, M. J., Hard, G. C., Hotz, K. J., Wang, C.-X., Williams, G. M., and Wilson, A. G. E. (1995). A study of the mechanism of butachlor-associated gastric neoplasms in Sprague-Dawley rats. Exp. Toxicol. Pathol. 47, 107–116. Thomas, G., and Williams, E. (1991). Evidence for and possible mechanisms of non-genotoxic carcinogenesis in the rodent thyroid. Mutat. Res. 248, 357–370. U.S. Environmental Protection Agency (EPA) (1994). Decision Document, Conditional Registration of the New Chemical Acetochlor, Office of Pesticide Programs, Washington, DC U.S. Environmental Protection Agency (EPA) (1995). Metolachlor Reregistration Eligibility Decision Document, Office of Pesticide Programs, Washington, DC U.S. Environmental Protection Agency (EPA) (1996). Proposed guidelines for carcinogen risk assessment: Notice. Fed. Reg. 17960–18011.

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U.S. Environmental Protection Agency (EPA) (1997). “Common Mechanism of Toxicity.” Presentation to the Scientific Advisory Panel, Washington, DC. U.S. Environmental Protection Agency (EPA) (1998a). Alachlor Reregistration Eligibility Decision Document, Office of Pesticide Programs, Washington, DC U.S. Environmental Protection Agency (EPA) (1998b). Propachlor Reregistration Eligibility Decision Document, Office of Pesticide Programs, Washington, DC. U.S. Environmental Protection Agency (EPA) (2001). The Grouping of a Series of Chloroacetanilide Pesticides Based on a Common Mechanism of Toxicity, Office of Pesticide Programs, Washington, DC. van de Sandt, J. J. M. (2000). “In Vitro Percutaneous Absorption Study with Propachlor in Ramrod® SC through Viable Human Skin Membranes.” Monsanto Unpublished Report. Von Schubert, H. (1979). Allergic contact dermatitis caused by propachlor. Dermatol. Monatsschr. 165, 495–498 (In German). Wester, R. C., Melendres, J., and Maibach, H. I. (1992). In vivo percutaneous absorption and skin decontamination of alachlor in rhesus monkey. J. Toxicol. Environ. Health 36, 1–12. Wester, R. C., Melendres, J. L., and Maibach, H. I. (1996). In vivo percutaneous absorption of acetochlor in the rhesus monkey: Dose-response and exposure risk assessment. Food Chem. Toxicol. 34, 979–983. Williams, R. (1971). Species variations in drug biotransformations. In “Fundamentals of Drug Metabolism and Drug Disposition” (B. Lau, H. Mandel, and E. Way, eds.), p. 187. Williams & Wilkins, Baltimore. Wilson, A. G. E., and Takei, A. S. (1999). Summary of toxicology studies with butachlor. J. Pestic. Sci. 25, 75–83. Wilson, A. G. E., Lau, H., Asbury, K. J., Thake, D. C., Heydens, W. F. (1995a). Mechanistic Basis for the Rat Specific Nasal Tumors Observed with Alachlor. Abstract, International Congress of Toxicology—VII, Seattle. Wilson, A. G. E., Lau, H., Asbury, K. J., and Heydens, W. F. (1995b). Metabolism of alachlor by human nasal tissue. Fundam. Appl. Toxicol. 15, 1398. Wilson, A. G. E., Thake, D. C., Heydens, W. R., Brewster, D. W., and Hotz, K. J. (1996). Mode of action of thyroid tumor formation in the male Long-Evans rat administered high doses of alachlor. Fundam. Appl. Toxicol. 33, 16–23. World Health Organization (WHO) (1993). “Propachlor.” Environmental Health Criteria Document 147 (prepared by L. Ivanova-Chemishanka). Zbinden, G. (1989). Hyperplastic and neoplastic responses of the thyroid gland in toxicological studies. Arch. Toxicol. 12, 98–106.

Chapter 83

Paraquat Edward A Lock School of Pharmacy and Biomolecular Sciences, Liverpool John Moores University, Byrom Street, Liverpool, L3 3AF, UK

Martin F Wilks Swiss Centre for Applied Human Toxicology, Basel, Switzerland

83.1  Identity, Properties and Use 83.1.1  Chemical Name Paraquat is 1,1-dimethyl-4,4-bipyridinium ion (IUPAC, CAS RN [4685-14-7]), also known as the 1,1-dimethyl4,4-bipyridyldiylium ion.

Dukatalon®, Opal®, Pathclear® (also includes simazine and aminotriazole), Preeglox®, Preglone®, Seccatuto®, Spray Seed®, Weedol®. Trade names of mixtures with urea herbicides include Dexuron®, Gramocil®, Gramonol®, Gramuron®, Tota-Col®.

83.1.4  Physical and Chemical Properties 83.1.2  Structure CH3

N+

N+ CH3 [2Cl–]

Paraquat dichloride

83.1.3  Synonyms The common name paraquat is in general use (BSI, E-ISO, ANSI, WSSA, JMAF), except in Germany. Paraquat is usually formulated as the dichloride salt (also known as methyl viologen) (CAS NR [1910-42-5]), the bis(methyl sulphate) salt (CAS NR [2074-50-2]) is no longer commercialised. Code designations for the material are PP148 (for the dichloride salt) and PP910 (for the bis(methyl sulphate) salt). Trade names for paraquat dichloride formulations include Crisquat®, Cyclone®, Dextrone X�®, Esgram�®, Efoxon�®, Goldquat�® 276, Gramoxone�®, Herbaxon�®, Katalon®, Osaquat Super®, Pilarxone®, R-Bix®, Speeder®, Starfire®, Sweep®, Total®, Weedless®. Mixtures of paraquat with diquat are sold under trade names including Actor®, Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

The molecular formula of the cation is C12H14N2 with a molecular weight of 186.3. The dichloride salt has the formula C12H14Cl2N2 and a molecular weight of 257.2. Paraquat dichloride forms colourless, hygroscopic crystals which decompose at 300°C. It is practically non-volatile with a vapour pressure of 0.1 mPa. It is very soluble in water (700 g/l at 20°C) and practically insoluble in most other organic solvents. It is stable in neutral and acidic media, but readily hydrolysed in alkaline media. Paraquat is photochemically decomposed by ultraviolet radiation in aqueous solution.

83.1.5  History, Formulations and Uses Paraquat was first described in 1882 by Weidel and Russo. In 1933, Michaelis and Hill discovered its redox properties and called the compound methyl viologen. The herbicidal properties of paraquat were first described by Brian et al. (1958) and it became commercially available in 1962. Paraquat is mainly formulated as an aqueous solution with surface-active agents. In some countries, a lowstrength granular formulation (also containing diquat) is available. Paraquat is a fast-acting, non-selective contact herbicide, absorbed by the foliage with some translocation 1771

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Table 83.1  Acute toxicity of paraquat to the rat (Data Expressed as mg paraquat ion/kg) Paraquat dichloride Sex

Strain

Route of administration

Median lethal dose (time studied)

Reference

Pure salts

F

NSa

po

112 ( 104–122), 150 (139–162), 141b (140–142) (14 days)

Clark et al. (1966)

Pure salt

F

NS

po

150 (110–173) (21 days)

Mehani (1972)

Formulation

M

Sprague Dawley

po

143 (123–166) (7 days)

Murray & Gibson (1971)

Formulation

M

Sherman

po

100 (87–117) (15 days)

Kimbrough & Gaines (1970)

Formulation

F

Sherman

po

110 (90–134) (15 days)

Kimbrough & Gaines (1970)

Formulation

M

Sprague Dawley

po

115 ( 90–150) (30 days)

Sharp et al. (1972)

Formulation

M

Wistar

po

95 ( 79–114) (30 days)

Sharp et al. (1972)

b

Pure salt

F

NS

ip

19 (16–21) 16 (14–19)

Clark et al. (1966)

Pure salt

F

NS

ip

16 (10–26) (21 days)

Mehani (1972)

Formulation

M

Sprague Dawley

iv

21 (19–25)

Sharp et al. (1972)

a

Not stated Dimethosulphate salt

b

in the xylem. It is used for broadspectrum control of broadleaved weeds and grasses in fruit orchards and plantations, and for inter-row weed control in many crops. It is also used for general weed control on non-crop land, as a defoliant on cotton and hops, for destruction of potato haulms, as a desiccant, and for control of aquatic weeds. Paraquat is rapidly deactivated upon contact with the soil and does not leach.

1972). Rabbits, however, do not show signs of respiratory distress, they stop eating and drinking and tend to die without overt toxicity, following oral dosing (Clark et al., 1966; Butler and Kleinerman, 1971). Rats and mice given doses above the median lethal dose, by ip or sc administration show signs of hyperexcitability, ataxia and convulsions and usually die within a few hours of dosing, indicative of an effect on the central nervous system (Clark et al., 1966; Bagetta et al., 1992). Following chronic exposure signs of toxicity are few, but may include respiratory effects.

83.2  Toxicity to Laboratory Animals Regulatory reviews on the toxicology of paraquat have been published (IPCS, 1984; JMPR, 2003) which contain unpublished proprietary information that has been submitted to support registration of paraquat. The following review is based on peer-reviewed published information but will refer to the proprietary information when relevant.

83.2.1  Signs of Toxicity Following a lethal dose of paraquat to rats mortality is first seen on days 2–5 after dosing but deaths can also occur around days 10–12 (Clark et al., 1966; Sharp et al., 1972; Smith and Rose, 1977), indicating there is considerable inter-individual animal response to the chemical. The major cause of death after a median lethal dose (MLD) is due to lung damage, the animals develop acute pulmonary oedema with signs of laboured respiration and ultimately die of respiratory failure (Clark et al., 1966; Kimbrough and Gaines, 1970; Sharp et al., 1972; Murray and Gibson,

83.2.2  Acute Toxicity The acute oral toxicity of paraquat to the rat is shown in Table 83.1. The MLD of pure paraquat dichloride expressed as the cation was about 150 mg/kg to female rats and ranged from 100–143 mg/kg in a number of different strains from a number of different laboratories. No sex difference in toxicity was seen and the toxicity was similar for the two different salts of paraquat (Table 83.1). Fasting rats prior to oral administration of paraquat made little difference to the toxicity, the 7 day MLD with 95% confidence limits were 143(123–166), 130(106–159) and 126(102– 156) mg paraquat ion/kg respectively for rats fasted for 0, 4 and 8 hr (Murray and Gibson, 1971). Mice are less sensitive than the rats to orally administered paraquat, while guinea pigs, cats, monkeys and rabbits are more susceptible (Table 83.2; JMPR, 2003). Paraquat was more toxic when given by the intraperitoneal (ip) or intravenous (iv) routes with an MLD of approximately 20 mg paraquat ion/kg (Table 83.1; JMPR,

Chapter | 83  Paraquat

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Table 83.2  Acute Toxicity of Paraquat to Laboratory Animals (Data Expressed as mg paraquat ion/kg) Species

Sex

Route of administration

Median lethal dose

Reference

Mouse

Fa

ip

30c

Ecker et al. (1976b)

c

F

30 (26.5–35.1)

Bus et al. (1976a)

Mouse

F

po

196c

Bus et al. (1975b)

Guinea pig

F

ip

3

Clark et al. (1966)

Guinea pig

Mb

po

30 (22–41) c

Clark et al. (1966)

Guinea pig

M&F

po

22 (15–33)

Murray and Gibson (1972)

Cat

F

po

35 (27–46)

Clark et al. (1966)

c

Monkey (Macaca fascicularis)

M&F

po

50

Rabbit

M

po

100c

Kuo & Nanikawa (1990)

Rabbit

M

po

50 (45–58)

Mehani (1972)

Rabbit

M

ip

25 (15–30)

Mehani (1972)

Rabbit

M

ip

18 (11–31)

McElligott (1972)

Dog

M

Dog

F

sc sc

Murray and Gibson (1972)

c

Nagata et al. (1992a)

c

Nagata et al. (1992a)

1.8 (1–6.1) 3.5 (2.4- 10.1)

a

Female Male c Paper refers to paraquat, not clear if salt or ion. b

Table 83.3  Acute Toxicity of Paraquat to Laboratory Animals Following Dermal Application or Inhalation Exposure (Data Expressed as mg Paraquat ion/kg). Species

Sex

Route of administration

Median lethal dose

Reference

Rat

M

Dermal

80 (60–96)

Kimbrough and Gaines (1970)

Rat

F

Dermal

90 (74–110)

Kimbrough and Gaines (1970)

Rabbit

M

Dermal

236 (collars removed)

Clark et al. (1966)

Rat

M&F

Inhalation

480 (with collars) 6 g/L. hr.

McElligott (1972) Gage (1968a)

2003), indicating that following oral dosing the compound is poorly absorbed from the gastrointestinal tract (see later). The guinea pig and dog (Nagata et al., 1992a) are also more sensitive to systemic administration with a MLD of 2–3 mg/kg (Table 83.2), reflecting poor or incomplete absorption of paraquat from the gastrointestinal tract after oral administration. Rabbits given a single iv dose of paraquat at 40 or 80 mg/kg died within 24 h, while they survived a single dose of 10 mg/kg iv, no lung lesions were seen at these doses (Ilett et al., 1974). The vehicle used to administer paraquat can influence lethality in mice. For example, paraquat was more toxic when given by the ip

or subcutaneous (sc) route in water than in isotonic saline, suggesting that the solvent may influence the absorption from the site of injection and hence the amount delivered to the lung (Drew and Gram, 1979). The dermal toxicity of paraquat has been studied in rabbits (Table 83.3; JMPR, 2003). The precise technique of application of paraquat to the skin, whether the site of application is open to the air or covered and whether the rabbits are prevented from grooming, affects the findings (Clark et al., 1966; McElligott, 1972). Rabbits fitted with restraining collars to reduce grooming the site of application, followed by decontamination of the skin and removal

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of the collars, showed glossitis, anorexia, weakness and loss of weight with some skin erythema followed by hyperkeratosis and desquamation at the higher doses, indicating that some oral ingestion had still occurred. This technique resulted in a MLD following a single application of 236 mg paraquat ion/kg. If however, the restraining collars were not removed then the erythema and desquamation was mild and the extent of glossitis and hence body weight loss was less. Under these conditions the MLD was found to be 480 mg ion/kg, the maximum dose possible to apply in a satisfactory manner (McElligott, 1972). Thus when compared to the systemic MLD of 18 mg ion/kg (Table 83.2) indicates that little of the applied dose has been absorbed through intact skin. Dermal exposure of rats to paraquat gave an MLD of 80–90 mg paraquat ion/kg (Table 83.3); however these authors (Kimbrough and Gaines, 1970) gave no information on the state of the skin after application, whether the site was occluded or free for the rats to groom. The absorption of paraquat across the skin has been reviewed by Smith, (1988) who concluded that paraquat is poorly absorbed by intact skin and raised technical concerns about the validity of the earlier dermal studies reported by Kimbrough and Gaines, (1970). Paraquat is not volatile, but following inhalation exposure to an aerosol is irritant to the respiratory tract (JMPR, 2003). At lethal concentrations under these conditions, death is usually delayed for several days and is due to respiratory failure. Following single exposures the MLD is a function of both the amount and duration of exposure, which in the rat is approximately 6 g/L. hr.(Table 83.3). Guinea pigs and male mice are of similar sensitivity to the rat, while female rats and rabbits are less sensitive. Dogs can tolerate a concentration-time product of 25 g/L. hr. without ill effects (Gage, 1968a). The toxicity is also a function of particle size, 3 m was the most lethal to the rat. Large particles do not reach the alveolar region and are less toxic. Under normal conditions of manufacture, handling and use inhalation exposure is not considered to be a hazard. Studies with rabbits have shown that the lung is susceptible to paraquat injury following intrabronchial deposition (Zavala and Rhodes, 1978) and inhalation exposure (Seidenfeld et al., 1978), although as mentioned above it is refractory following oral or intraperitoneal administration. Local instillation of paraquat in the lungs of rats will also produce local injury and fibrosis (Kimbrough and Gaines, 1970; Wyatt et al., 1981).

83.2.3  Irritation and Sensitization Paraquat is a skin and eye irritant but is not a skin sensitiser (Bainova, 1969; JMPR, 2003). As discussed earlier skin irritation has been reported in rabbits when the area of application is occluded (Clark et al., 1966; McElligott, 1972), resulting in local erythema followed by hyperkeratosis and desquamation.

Hayes’ Handbook of Pesticide Toxicology

Instillation of a 0.29% aqueous solution of paraquat into the rabbit eye produced no effect; however more concentrated solutions produced inflammation of the conjunctiva and nictitating membrane. This response developed gradually over 12 h and lasted for 48–96 h (Clark et al., 1966). Instillation of higher concentrations of paraquat (3–48 mg contained in 0.2 ml of water into the rabbit eye produced a dose-related increase in ocular injury with doses of 48 mg (about 16 mg/kg) and above producing fatalities (Sinow and Wei, 1973). These findings indicate that absorption of paraquat from the eye is similar to that following systemic administration.

83.2.4  Subchronic Toxicity Daily administration of paraquat in the diet of rats at an inclusion rate of 100 ppm (about 5 mg ion/kg/day) was tolerated for several months. However, if increased to 250 ppm (about 12.5 mg ion/kg/day) the rats became ill and died within 27 to 57 days. Females appeared to be more susceptible than males, the primary target organ for toxicity being the lungs (Clark et al., 1966). A number of other studies have shown that moderate daily doses of paraquat can be tolerated for up to 2 years (JMPR, 2003). Rats fed a dietary concentration of 25 ppm (about 1.25 mg ion/kg/day) for 2 years showed mild ocular lesions. Mice fed paraquat for 2 years showed a no adverse effects at 12.5 mg/kg (equivalent to 1.88 mg of paraquat ion/kg bw per day) on the basis of decreased body-weight gain in females and renal changes in males at the next highest dietary concentration. Dogs tolerated 50 ppm (about 0.9 mg/kg/day) for 2 years (Howe and Wright, 1965), although in a later 1 year study the no effect level was 15 mg/kg on the basis of erythema of the tongue at 30 mg/kg in males, elevated alkaline phosphatase in both sexes, and histopathological changes in the lung at  30 mg/kg, this is equal to 0.45 mg of ion/kg/day (JMPR, 2003). Rats given paraquat in drinking water at 1.3 mg/L or 2.6 mg/L for 2 years showed some mortality and histological changes in the lung at the highest dose, only minimal changes to the lung were seen at the lower dose (Bainova and Vulcheva, 1977). The MLD for paraquat fed in the diet for 90 days has been determined. Groups of female rats were fed 300, 400, 500, 600 and 700 ppm paraquat and their food consumption recorded at intervals to enable the dose in mg/kg/ day to be calculated. After 90 days the surviving rats were held for 2 weeks to allow time for any delayed deaths. The MLD was 21 mg/kg/day, giving a subchronic toxicity factor of 5.2 (ratio of acute to subchronic MLD’s), indicating that paraquat has a moderate cumulative toxicity in this species (Kimbrough and Gaines, 1970). Rabbits given paraquat ip at 10 or 20 mg/kg at 48 h intervals showed marked signs of toxicity with a high mortality following 3 to 5 doses, there was little evidence of lung damage and it is likely that the animals died from multiorgan failure (Butler and Kleinerman, 1971). Rabbits can

Chapter | 83  Paraquat

however tolerate 3 mg/kg day ip for up to 14 days, but when increased to 6 mg/kg/day significant mortality was seen (Hassan et al., 1989). Daily oral dosing at 11 mg/kg/day to male rabbits for 30 days produced few signs of toxicity, with only one animal showing lung damage (Dikshith et al., 1979). Subchronic exposure following dermal application has been examined in rabbits. The mortality observed with repeated daily applications beneath an occlusive dressing gave a MLD of 6.24 (4.6–8.5)mg ion/kg/day of paraquat over 20 days (McElligott, 1972). At the higher doses the skin was reddened and sloughing with local oedema was observed, while at the lowest dose some scab formation was seen after about 7 days of application. Systemic effects at post-mortem included renal tubular necrosis, focal hepatocellular necrosis and pulmonary congestion. Studies were also conducted where the skin was not occluded, rabbits were fitted with collars and these were either removed after decontamination of the skin or at the end of the observation period. The MLD for 20 days exposure was between 7.25 and 14.5 mg paraquat ion/kg/day for the animals where the collars were removed after decontamination and at least 24 mg ion/kg/day for those where the collars were left on all the time. The rabbits showed marked signs of salivation, which was associated with glossitis and ulceration of the tongue. The animals refused to eat and death occurred in a state of cachexia; this effect was less marked at the lower doses when the collars were kept in place all the time (McElligott, 1972). When the “Gramoxone” formulation of paraquat was diluted to spray strength and applied to the skin of rabbits for 20 days (2.4 mg ion/kg/day) no clinical signs of toxicity or pathological changes were seen (McElligott, 1972). A more recent study in rabbits given paraquat for 6 h/day for 21 days reported a no effect level of 1.5 mg/kg/day of paraquat dichloride (JMPR, 2003). Daily subcutaneous dosing of paraquat to dogs for 4 weeks, resulted in some animals being terminated at the top dose of 0.495 mg/kg/day, while at the other doses of 0.165 and 0.055 mg/kg/day all animals appeared well (Nagata et al., 1992b). Histopathology of the lungs showed proliferation of alveolar lining cells and some fibrosis at the top dose and pulmonary changes (thickening of the alveolar wall and pleura) at all doses. The 28 day MLD from this study was about 0.5 mg/kg/day. More recently studies in dogs exposed to paraquat for 6 weeks, 13 weeks and 1 year have been reported (JMPR, 2003). Repeated exposure of rats to paraquat by inhalation at 0.4 g/L for 6 hr per day for 15 days, over 3 weeks, led to intermittent respiratory problems after about 4 exposures. At post-mortem after 15 exposures the animals showed marginal paraquat-related pathology to the lungs. While, exposure to 0.1 g/L for 15 daily, 6 hr periods showed no signs of toxicity or pathology in the lungs (Gage, 1968a). Rats exposed to 0.003 g/L for 6 hr a day for 5 days per week for 2 months, put on body weight, remained in good

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condition and showed no histopathological evidence of lung damage. Bainova et al. (1972) exposed rats to a paraquat aerosol at 1.1 or 0.05 mg/L for 6 hr /day for 4.5 months and found evidence of lung damage at the higher dose, with little effect at the lower dose. Seidenfeld et al. (1978) exposed rabbits by inhalation to paraquat by an ultrasonic nebuliser (mean particle size 4 m) at a concentration of 0.1 mg/ml for 2 hr/day for 5 days per week for 3 months and found no lung damage, however rabbits exposed to 2 mg/ml for 2 hr/day could only tolerate 3 exposures and developed a reduced arterial oxygen tension and specific compliance which was associated with marked lung injury. Mice exposed to paraquat in the diet at concentrations of 10, 30, 100 and 300 mg/kg/day for 13 weeks, showed lung injury at the top dose with a no adverse effect level of 100 mg/kg/day (JMPR, 2003). Overall these studies show that following acute and chronic exposure the primary target organ for toxicity is the lung, with deaths from lung damage frequently taking many days to occur following a single dose. The rabbit is unusual in that it does not readily develop a lung lesion following oral or parenteral exposure, but if instilled into the lung or exposed via inhalation, lung injury ensues. Renal functional impairment with some renal tubular necrosis is the other major organ affected. Dose levels of paraquat that do not cause lung damage in laboratory animals following acute and chronic exposure have been clearly established.

83.2.5  Mutagenic and Carcinogenic Potential Paraquat is not carcinogenic in either rats or mice (JMPR, 2003). Activity seen in some short-term assays for mutagenesis is associated with cytotoxicity, and believed to arise as a consequence of the redox cycling ability of paraquat, leading to superoxide anion formation. Paraquat has minimal to no genotoxic activity when evaluated in a wide range of in vitro and in vivo test systems. Many groups have reported the absence of an effect while others have reported weakly positive effects (IPCS, 1984; Ribas et al., 1995; Dabney, 1995, JMPR, 2003). These later effects were usually associated with high cytotoxicity or mortality and are believed to arise as a consequence of the redox cycling ability of paraquat. It is known that DNA damage frequently occurs when cells are exposed to oxidative stress (Brawn and Fridovich, 1981; Repine et al., 1981). Paraquat-mediated effects on DNA have been reported in bacteria (Moody and Hassan, 1982; Yonei et al., 1986), Chinese hamster cells (Sofuni et al., 1985; Nicotera et al., 1985;Takana and Amano, 1989), isolated alveolar macrophages and epithelial type II cells (Dusinska et al., 1998) and in a few cases in cells from treated mice (He and Yasumoto, 1994; Rios et al., 1995) these responses are all considered to be secondary to superoxide anion generation.

1776

Studies with cultured mammalian cells have shown that paraquat inhibits DNA synthesis leading to the arrest of the cells in S-phase (Yamagami et al., 1994; Tomita, 1996). This effect occurs prior to the onset of cytotoxicity and is thought to be part of a cascade of events initiated by the production of oxygen free radicals by the redox cycling of paraquat. These findings have been extended to rat lung cells exposed to paraquat in vivo which also showed S-phase arrest at early times after dosing. Prior treatment of the rats with a diet enriched in sodium tungstate, an inhibitor of xanthine oxidase, to reduce the production of free radicals, prevented the S-phase arrest produced by paraquat (Matsubara et al. 1996) and reduced mortality (Kitazawa et al., 1991). Once inside a cell, paraquat can redox cycle producing oxygen free radicals that can cause cell cycle arrest and inhibit DNA synthesis. These findings are consistent with early studies showing that paraquat reduces DNA synthesis at early times after dosing (Van Osten and Gibson, 1975; Smith and Rose, 1977). Paraquat has been evaluated for its carcinogenic potential in both rats and mice and it was concluded that at all doses up to the maximum tolerated dose, paraquat did not result in a compound related increase in tumour incidence (Bainova and Vulcheva, 1977; JMPR, 2003).

83.2.6  Effects on Reproduction, Embryotoxicity and Teratogenicity Paraquat has no effect on fertility, is not teratogenic and only produces fetotoxicity at doses that are maternally toxic. The main finding in multigeneration studies was lung damage (JMPR, 2003). Paraquat does not readily cross the placenta and enter the embryo of mice when given either orally or by ip administration (Bus et al., 1975). In contrast, paraquat appears to readily cross the placenta of rats, being detected in fetuses within 30 min of an iv injection to 20 day pregnant rats (Ingebrigtsen et al., 1984). A three-generation reproduction study in rats maintained on dietary levels of paraquat of 30 or 100 ppm showed no effect on food intake, fertility, fecundity, neonatal morbidity or mortality. No teratogenesis or other changes in gross or histological morphology were seen, except for a slight increase in the incidence of renal hydropic degeneration in the 3–4 week old young receiving 100 ppm (about 10 mg/kg/day). Pregnant and young animals did not appear to be more susceptible than adults (JMPR, 2003). A two-generation reproduction study in mice maintained on dietary levels of 45, 90 or 125 ppm showed no effects on age to parturition, number born or abnormalities in the pups in the first generation following 45 or 90 ppm. However, at 125 ppm an increase in mortality was seen in the dams and pups during the first few weeks of life (Dial and Dial, 1987). The second generation mice were more resistant to the effects of paraquat the only effect being an increase in the age of the mothers at second

Hayes’ Handbook of Pesticide Toxicology

parturition on the highest dose of paraquat (Dial and Dial, 1987). Subsequent studies to explore the basis for the high mortality in the first generation dams and pups exposed to 125 ppm paraquat in the diet, showed that they almost certainly died from lung damage. This only occurred in pups exposed pre-natally via the placenta, not in pups exposed post-natally (Dial and Dial, 1989). Bus and Gibson (1975) also reported that paraquat given to mice in their drinking water at either 50 or 100 ppm from day 8 of gestation and to the young until 42 days of age increased pup mortality at 100 ppm but not 50 ppm. The lungs of mice killed 42 days after 100 ppm showed extensive alveolar consolidation and collapse, supporting the view that the deaths at this dose were probably due to lung damage. No dominant lethal effects were seen in mice exposed to paraquat at oral doses up to 4 mg/kg/day for 5 days (Anderson et al., 1976). High doses of paraquat injected ip into pregnant rats or mice on various days of gestation can produce significant maternal toxicity (Bus et al., 1975; Khera and Whitta, 1970). Examination of the foetuses of mice exposed to 1.67 or 3.35 mg/kg ip or 20 mg/kg po daily on days 8–16 of gestation induced no teratogenic effects, although a slight increase in non-ossification of the sternbrae was seen (Bus et al., 1975). With the discovery that paraquat can damage neurons in the substantia nigra of the mouse brain (see Section on Central Nervous System) a small number of studies have examined whether prenatal administration of paraquat or the pesticide maneb would have adverse effects on the development of the nigrostriatal doperminergic system. C57Bl/6 J mice were treated on gestational days 10–17 with saline or the pesticides maneb (1 mg/kg) or paraquat (0.3 mg/kg). None of the prenatal conditions caused any adverse effect on gestation, parturition, litter survival or growth at any point, nor was there any morphological or behavioral teratogenicity associated with any prenatal exposure. When the offspring were evaluated in adulthood, there were no significant effects of prenatal maneb or paraquat on locomotor activity. Subsequently, offspring were treated for 8 consecutive days with saline, maneb (30 mg/kg), or paraquat (5 mg/kg). One week after the last exposure, only males exposed to prenatal maneb and adulthood paraquat showed significant reductions in locomotor activity (95%),changes in striatal neurochemistry and striatal neuron cell loss. Pre-natal exposure to paraquat followed by paraquat in adulthood did not enhance neuronal cell loss in the substantia nigra (Barlow et al., 2004; Cory-Slechta et al., 2005; Barlow et al., 2007).

83.2.7  Pathology of the Lung The toxic effects of paraquat were first described by Clark et al. (1966) who reported that the histological effects of paraquat in rats, mice and dogs are similar. The lung, liver, kidney and thymus were affected; the lung being the major target. The effect of paraquat in the cynomologus monkey

Chapter | 83  Paraquat

and Gage, (1966) in rats. Following a single oral dose of 4, 6 or 50 mg/kg [14C-methyl] paraquat dichloride most of the radioactivity was excreted within 48 h. Occasionally some appeared in the faeces 3 and 4 days after dosing at the higher doses, with small amounts also in the urine. Between 6 and 14% of the dose was excreted in the urine over 48 h when given as the dichloride salt, and 16–23% when given as the dimethylthiosulphate salt, the remainder being in the faeces. In contrast, when paraquat as either salt was given sc then the bulk of the radioactivity appeared in the urine within 24 h of dosing, showing that paraquat is poorly absorbed across the gastrointestinal tract of the rat. Subsequent studies have extended and essentially confirmed these findings (Molnar and Hayes, 1971; Murray and Gibson, 1974; Lock and Ishmael, 1979; Chui et al., 1988; JMPR, 2003). The concentration of paraquat in the plasma following an oral dose to the rat is determined largely by the amount of paraquat present in the small intestine (Smith et al., 1974). Studies in the dog using tracer doses (129 g/kg) of [14C-methyl]paraquat support this, as peak plasma concentrations following oral dosing were observed at 75–90 min (Figure 83.1), with about 46–66% of the dose absorbed, as judged by the amount excreted in the urine at 6 h (Davies et al., 1977). Thus, the dog absorbs a greater percentage of an orally administered dose of paraquat than the rat, which is consistent with the greater susceptibility of the dog to paraquat by this route of administration. Pre-treatment of dogs with a drug that will block gastric emptying, delayed the peak plasma concentration by 3 to 6 h, indicating that the stomach is not the major site of absorption (Bennett et al., 1976). These data in both rats and dogs indicate that the absorption of paraquat from the gastrointestinal tract occurs somewhere beyond the stomach. It is assumed this is similar for humans but there is limited evidence to support this. Based on the cationic nature of paraquat, it would not be expected to readily cross cellular membranes; it seems unlikely that simple diffusion would explain the rapid but incomplete absorption seen in the rat and dog.

25

1.6 Dog

20

Rat 1.2

15 0.8 10 0.4

5 0

Rat Plasma paraquat (ng/ml)

Dog Plasma paraquat (ng/ml)

is similar to that in rats (Murray and Gibson, 1972). In contrast, as mentioned previously, rabbits do not develop lung lesions following acute oral or ip administration (Butler and Kleinerman, 1971; Mehani, 1972; Ilett et al., 1974; Zavala and Rhodes, 1978). There is one report of daily administration of paraquat in the drinking water to rabbits over several days leading to lung damage that resembles that seen in rats (Restuccia et al., 1974). Inhalation exposure to paraquat produces lung damage in the rabbit (Seidenfeld et al., 1978). The hamster responds in a similar way being refractory to single sc dose of paraquat, but lung fibrosis is produced by repeated sc injections (Butler, 1975). The most extensive studies on the pathogenesis of lung damage produced by paraquat have been conducted in rats. The time course of development of the injury in rats given a single MLD ip was reported by Vijeyaratnam and Corrin (1971) and Smith and Heath, (1974a). Damage to the type I and II alveolar epithelial cells was seen within a day of dosing. This damage was more marked by days 2–4 with large areas of the alveolar epithelium being completely lost, alveolar oedema developed and in some areas haemorrhage into the air spaces occurred. At this time there was extensive infiltration of inflammatory cells into the alveolar interstitium, air spaces and perivascular areas, although the alveolar endothelial capillaries were mainly spared. The animals died as a consequence of severe anoxia usually within the first few days after dosing and this has been confirmed by others (Clark et al., 1966; Sharp et al., 1972; Smith and Rose, 1977). This phase has been called the destructive phase (Smith and Heath, 1976). Similar early pathological changes have been reported by Kimbrough and Gaines, (1970); Brooks, (1971); Modee et al. (1972); Wasan and McElligott, (1972); Smith et al. (1974); Sykes et al. (1977) and Smith and Heath, (1976). Some rats that survive for up to 10–12 days after dosing develop an extensive hypercellular lesion in the lung which is dominated by proliferation of fibroblasts. This phase of the lesion is called the proliferative phase and is characterised by attempts by the epithelium to regenerate and restore normal architecture of the alveolar epithelium (Kimbrough and Gaines, 1970; Vijeyaratnam and Corrin, 1971; Smith and Heath, 1974a). The findings in these animals are typically extensive intraalveolar and inter-alveolar fibrosis, which in association with residual oedema reduces gaseous exchange which results in death from anoxia. It appears that the initial damage to the alveolar epithelium, produced by paraquat, is the primary event in the development of the lung injury, with the proliferative fibrosis being a consequence of the extensive damage produced. For a more detailed review on pulmonary injury see the review by Smith and Heath, (1976).

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0 0

1

2

3

4

5

6

Time (h)

83.2.8  Absorption The first studies on the absorption and excretion of paraquat from the gastrointestinal tract were conducted by Daniel

Figure 83.1  Plasma levels of paraquat in the rat and dog following a single non-toxic oral dose. The dog was given a total dose of 1.03 mg of paraquat, while the rats were dosed at 0.038 mg/kg. Data adapted from Davies et al. (1987) and Chui et al. (1988).

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Plasma paraquat (µg/ml)

5 Rat

4

Monkey

3 2 1 0 0

6

12

18

Time (h) Figure 83.2  Plasma levels of paraquat in the rat and monkey following a single toxic oral dose. The rats were given 126 mg/kg paraquat while the monkeys received 50 mg/kg. Data adapted from Murray and Gibson, (1974).

Studies in vitro with isolated mucosa from a number of different regions of the rat gastrointestinal tract (Steffen and Konder, 1979) have confirmed that the jejunum and ileum have the greatest capacity to transport paraquat from the lumen into the bloodstream and also showed that a component of the transport is facilitated (Heylings, 1991). Following oral administration of paraquat to rats, the peak plasma concentration is seen between 30–60 min (Figure 83.1 and 83.2) following either a tracer dose (Chui et al., 1988) or a toxic dose (Murray and Gibson, 1974). This profile is similar to that seen in the dog (Figure 83.1) (Davies et al., 1977). The peak plasma concentration in the monkey and guinea pig occurs within the first hr (Figure 83.2) and 30 min respectively following a toxic oral dose (Murray and Gibson, 1974). Overall, these studies indicate that paraquat is rapidly but incompletely absorbed from the gastrointestinal tract of laboratory animals and humans (see later), with peak plasma concentrations occurring within 30–90 minutes. Paraquat is poorly absorbed across human skin in vitro, human skin being less permeable to paraquat than the skin of rats, rabbits or guinea pigs (Walker et al., 1983). Application of a low dose of [14C] paraquat (150 nmol/kg) in acetone to rat skin resulted in a peak blood levels about 1 hr after dosing and a total of 3.5% of the dose absorbed (Chui et al., 1988). It should be pointed out that a occlusive dressing was applied in these studies which has previously been shown to greatly enhance the percutaneous absorption of paraquat in animals (McElligott, 1972). Overall, these studies plus those of Hoffer et al. (1989) in rabbits, indicate that paraquat is poorly absorbed across the intact skin of laboratory animals.

83.2.9  Distribution In the rat, after a lethal oral dose, the plasma paraquat concentration remained relatively constant after the initial

peak for up to 32 h (Murray and Gibson, 1974; Rose et al., 1976a), during this time the concentration in the lung rose progressively to several times that found in the plasma. In no other organ, apart from the kidney, the major organ for the excretion of paraquat, was a time-dependent accumulation of paraquat detected (Murray and Gibson, 1974; Rose et al., 1976a). These findings, plus the earlier observation of Sharp et al. (1972) who administered paraquat iv and showed that paraquat was retained in the lung with a halflife of 50 h, provided the key evidence showing that those organs that had the highest concentration of paraquat were those that were susceptible to injury, namely the lung and kidney. Many other groups have subsequently examined the pharmacokinetics and elimination of paraquat in the rat (Maling et al., 1978; Chui et al., 1988; Dey et al., 1990; JMPR, 2003), dog (Hawksworth et al., 1981; Giri et al., 1982; Pond et al., 1993) rabbit (Ilett et al., 1974; Yonemitsu, 1986; Yu et al., 1994) and mouse (Drew and Gram, 1979; JMPR, 2003). The distribution of paraquat in the body is best described by a three-compartment model, with input to and removal from the central plasma compartment. Simulations of plasma concentrations in the peripheral compartments show there is a compartment with rapid uptake and removal of paraquat, which was assumed to be the highly vascular tissues such as the kidney, and a slow uptake compartment reaching a maximum about 4–5 hr after iv dosing , which may be the lung (Hawksworth et al., 1981). Using lung slices, Rose et al. (1974a) first described the time-dependent accumulation of paraquat into lung tissue. This process was shown to be energy-dependent in that it could be inhibited by the addition of the metabolic inhibitors cyanide plus iodoacetate to the incubation medium. The accumulation of paraquat into rat lung was shown to obey saturation kinetics with an apparent Km of 70 M and a Vmax of 300 nmol/h/g wet weight of lung slice (Table 83.4) (Rose et al., 1974a). Other aspects of the accumulation of paraquat into the lung will be discussed in more detail later. Hawksworth et al. (1981) also showed that early onset of renal failure markedly affected the concentration of paraquat in the peripheral compartments, suggesting that any reduction in renal excretion of paraquat may allow more of the chemical to be transported into the lung. The distribution in the rabbit, which is refractory to lung damage following a single systemic dose, showed the organs with highest concentration of paraquat were the lung and kidney at 6 and 24 after dosing but the concentration in rabbit lung appeared to decline more rapidly than from rat lung (Ilett et al., 1974). Whole body autoradiography studies have provided valuable information on the tissue distribution of paraquat, early studies by Litchfield et al. (1973) in mice given iv [14C methyl]-paraquat, showed retention in the lung. A more detailed study using [3H-methyl]- paraquat and thin tissue sections revealed localisation of radioactivity at all time intervals after dosing in the lung, choroid plexus

Chapter | 83  Paraquat

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Table 83.4  Kinetic Constants for the Accumulation of Paraquat and Putrescine by Rat and Human Lung Slices or Isolated Alveolar Type Ii Cells Species/Tissue

Rat-lung slice

Paraquat accumulation

Putrescine accumulation

Km (M)

Vmaxa

  70

300

Rose et al. (1974a)

210

710

Ross & Krieger (1981)

119

636

Km (M)

Vmaxa

8

480

Karl & Friedman (1983)

7

330

Smith & Wyatt (1981)

31

870

Nemery et al. (1987)

12–18

Human-lung slice

  40

300

244

370

Cultured rat-type II cells

O’Sullivan et al. (1991)

13.5

720

Hardwick et al. (1990)

13.1

723

Smith et al. (1982) Rose et al. (1974a)

7

376

Hoet et al. (1994)

2–11

99–249

Brooke Taylor et al. (1983)

7

414

Hoet et al. (1993)

5

18

Lewis (1989)

8–14

58

Richards et al. (1987)

  29

Van der Wal et al. (1990)

  64 Suspensions of rat-type II cells

  88

  29

Cultured human-type II cells a b

Reference

15

128

Oreffo et al. (1991)

2.5

34

Chen et al. (1992)

6–8

12–14

Hoet et al. (1994)

Vmax in lung slices expressed as nmol/h/g wet weight of slice. Vmax Alveolar type II cells expressed as pmol/h/M DNA.

(see also Section 7.2.20), muscle and melanin in addition to excretory pathways such as the proximal tubules of the kidney (see also Section 7.2.19), urine, liver, gall bladder and intestinal contents of the mouse (Waddell and Marlowe, 1980). Radioactivity in the lungs appeared to be higher in certain areas and higher cellular resolution autoradiography revealed that the radioactivity was confined to alveolar type II cells, which are one of the major target cells for paraquat toxicity. In these studies it was essential to keep the tissue frozen at all times to prevent diffusion of paraquat which is highly polar. An association of paraquat with melanin has been demonstrated and this is probably due to an ionic interaction (Larsson et al., 1977, 1978; Lindquist et al., 1988). Immunohistochemical approaches utilising specific antibodies to paraquat have shown immunoreactive material localised primarily in bronchiolar epithelial cells and walls of blood vessels in the lungs of rats, 3 h to 10 days after an iv dose. Other studies have localised

immunoreactive material in the intestine, liver, kidneys and in the brain to capillary walls and glial cells but not neurones (see also Section 7.2.20), after paraquat administration (Nagao et al., 1990; 1991; 1993).

83.2.10  Metabolism Paraquat is very poorly metabolised with the bulk of the administered dose being excreted unchanged in the urine and faeces. Daniel and Gage, (1966) compared the colourimetric assay for paraquat with that found by radiochemical detection on the urine and faeces of rats dosed with paraquat and demonstrated that there was very close agreement. Chromatography of the urine and lung tissue from rats treated with paraquat also showed no evidence of biotransformation (Hughes et al., 1973; Murray and Gibson, 1974; Rose et al., 1974a). No radioactivity was excreted in expired air following paraquat administration to rats, indicating that

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1780

it did not undergo metabolism to CO2 (Murray and Gibson, 1974). Incubation of paraquat with rat caecal contents for up to 24 h showed up to a 50% loss, indicating microbial metabolism, the loss was not seen when the contents of the caecum were heat treated (Daniel and Gage, 1966). However, in vivo studies in rats, guinea pigs and dogs showed little evidence of biotransformation, indicating that the in vitro studies had over predicted the likely metabolism (Summers, 1980). The overriding weight of evidence is that metabolism does not contribute to the toxicity of paraquat.

83.2.11  Excretion Elimination of paraquat from the body is almost exclusively via the kidneys. The renal clearance of paraquat is greater than that of creatinine in the rat (Lock, 1979; Chan et al., 1997); dog (Hawksworth et al., 1981); sheep (Webb, 1983); monkey (Purser and Rose, 1979) and humans (Bismuth et al., 1982); see later for more detailed discussion on humans. Thus paraquat is actively secreted by the kidney. Renal tubular secretion was completely inhibited by N1-methylnicotinamide suggesting that paraquat is secreted via a cationic transport system (Hawksworth et al., 1981). The transport mechanisms for organic cations in renal proximal tubular cells is not fully understood. Two membrane proteins, organic cation transporter 1 (OCT1) (Grundemann et al., 1994) and organic cation transporter 2 (OCT2) (Okuda et al., 1996) have been isolated from rat kidney. OCT1 located on the basolateral membrane, will transport tetraethylammonium, and this can be inhibited by other organic cations such as quinine. OCT2, which is predominantly expressed in the kidney, stimulates the uptake of tetraethylammonium and this can be markedly inhibited by cimetidine. Studies using freshly isolated renal proximal tubules and renal cell lines have shown that paraquat is transported across the basolateral membrane (from the blood stream into the renal tubular epithelial cell) using an OCT system (Groves et al., 1995; Chan et al., 1996 a and b, 1997, 1998). The transport of paraquat can be blocked by the addition of the divalent cation quinine, cimetidine and to a lesser extent tetraethylammonium (Chan et al., 1996b) suggesting that paraquat may be transported by both

transport systems. Recent studies have transfected human OCT,1, 2 and 3 into human kidney cells and examined the transport of paraquat, showing it is a good substrate for OCT2, limited uptake was seen with OCT1 and no uptake via OCT3 (Chen et al., 2007). Exit across the apical membrane into the tubular lumen is also an active process, and evidence suggests that there are two cation transport systems, an electroneutral organic cation/H exchange (Sokol et al., 1988) and P-glycoprotein (Dutt et al., 1992). Studies with rabbit brush-border membrane vesicles have shown that paraquat is a substrate for the cation/H exchange transporter, and further that it can inhibit the transport of other monovalent cations such as tetraethylammonium (Wright and Wunz, 1995). Recently two transporters for organic cation’s have been cloned and identified in human kidney as members of the multidrug and toxin extrusion (MATE) family and named MATE-1 and MATE-2 K (Otsuka et al., 2005; Masuda et al., 2006; Zhang et al., 2007). These are the electroneutral organic cation/H exchange transporters mentioned above. Human MATE-1 has been transfected into human kidney cells and shown to transport paraquat. Thus it is now clear that paraquat can enter a renal cell via OCT2 and to a lesser extent OCT1 and then be transport out of the cell by MATE-1 (Figure 83.3). Whether MATE-2 K can transport paraquat is not known at this time. In the rat in vivo, the fractional excretion of paraquat decreased from 2.1 at a plasma concentration of about 0.4 nmol/ml to 1.2 at a plasma concentration of 21 nmol/ml, demonstrating that the excretion of paraquat is greater than the glomerular filtration rate and that the process is saturable (Chan et al., 1997). Thus, at low plasma concentrations paraquat will be readily cleared from the body, however at higher plasma concentrations the system will become saturated and less paraquat will be cleared. At toxic doses it is well established that paraquat can cause renal functional impairment. In rats, given 126 mg ion/kg, po (Lock, 1979) and mice given 50 mg ion/kg, iv (Ecker et al., 1975b) renal impairment was observed 17–24 h after dosing. In the cynomologus monkey given 85 mg ion/kg, po the decline in renal clearance was seen 12 h after dosing the first time examined (Purser and Rose, 1979). In dogs given 20 mg ion/kg,

Renal tubule cell Cimetidine quinine

PQ2+

MATE-1

OCT1

PQ2+

PQ2+ PQ2+

P-glycoprotein

OCT2 pH 7.2 pH 7.4 Basolateral membrane

pH 6.7 Apical membrane

PQ2+

Figure 83.3  Mechanism of paraquat transport across renal tubular cells.  A schematic representation of the proposed transport systems for paraquat across renal tubular cells. The transporters are organic cation transporters 1 and 2 (OCT 1 & OCT 2) at the basolateral membrane and P-glycoprotein and multidrug and toxic extrusion transporter (MATE-1) a cation/H exchange system at the brush border membrane. Adapted from Chan et al. (1998). Reproduced with permission from Pharmacol. Ther.

Chapter | 83  Paraquat

iv (Hawksworth et al., 1981) renal impairment was observed as early as 2.5 hr after dosing. An early report on the renal handling of paraquat by the dog, suggested that paraquat was reabsorbed by the proximal tubules, this study was conducted at high plasma concentrations (54–810 nmol/ml) where the transport system will have been saturated and function impairment will almost certainly will have occurred (Ferguson, 1973). The weight of evidence strongly supports the view that paraquat is actively secreted by the kidney of laboratory animals and humans (see later). The implication of impairment of renal excretion is that more paraquat is available in the plasma, to accumulate into the lung. Whole body autoradiography has shown that paraquat was present in the gall bladder of mice, indicating some biliary excretion (Waddell and Marlowe, 1980). The extent of biliary excretion of paraquat was 5% when dosed to bile cannulated rats, rabbits or guinea pigs and measured over a three hour period (Hughes et al., 1973). The bulk of the dose appeared unchanged in the urine, these authors suggesting that the molecular weight of paraquat at 186 was below the minimal molecular weight of about 500 for chemicals that are excreted in bile. Radioactivity from paraquat was also detected in the bile of dogs given a single iv dose, indicating some biliary excretion in this species (Giri et al., 1982).

83.2.12  Accumulation of paraquat   into the lung The original discovery of an energy-dependent accumulation of paraquat into rat lung tissue (Rose et al., 1974a), lead to studies to look for this transport system in the lung of other species, including human. The accumulation of paraquat by slices of lung from a number of species was reported by Rose et al. (1976a). The apparent kinetic constants for the uptake process were very similar for all species examined except the rabbit. Slices of rabbit had a very high affinity, but low capacity, to accumulate paraquat which is consistent with the in vivo findings that show that following oral or parenteral administration of paraquat the rabbit does not develop a lung lesion. For the rat the derived Km was 70 M with a Vmax of 300 nmol/h/g wet weight of lung (Table 83.4). The kinetic constants for rat and human lung were very similar, suggesting that the rat lung was a good surrogate for studying paraquat uptake into human lung (Rose et al., 1976a). The kinetics of accumulation of paraquat into human lung slices has been confirmed by others, the Vmax being similar at 370 nmol/h/g wet weight while the Km was lower at 244 M (Hoet et al., 1994). Considerable inter-individual variation is seen in paraquat accumulation into human lung slices (Brooke-Taylor et al., 1983) which may either reflect individual variability or more likely the state of the tissue and delay between removal of the tissue and analysis of paraquat transport. Table 83.4 summarises the available

1781

data on the transport kinetics for paraquat in rat and human lung tissue. These observations, coupled with the finding that paraquat is not metabolised by the lung nor covalently bound to any degree (Ilett et al., 1974; Forman et al., 1982; Sullivan and Montgomery, 1983), suggests that this accumulation is mediated through binding to, and subsequent translocation into lung cells by a carrier-mediated system. It is interesting that paraquat which can enter renal cells via OCT2 (see earlier) does not appear to use this system in the lung. OCT2 is present in human and rodent lung being expressed in the respiratory epithelium of trachea and bronchi (Koepsell et al., 2007) but it role in paraquat transport if any is not known. The finding that paraquat was actively transported into lung slices lead to a search for chemicals that might inhibit this process (Lock et al., 1976; Maling et al., 1978; Smith et al., 1981; Ross and Krieger, 1981; Dunbar et al., 1988) and hence provide protection against paraquat-induced lung toxicity. A number of chemicals were identified that could block paraquat uptake into lung slices but none of these were effective in the whole animal (see later under treatment of poisoning). Studies were also undertaken to try and identify the endogenous chemicals for this transport system. A wide range of chemicals were examined and a number of naturally occurring amines were identified as the most effect inhibitors of paraquat accumulation into slices of rat lung, and which themselves act as substrates. These amines include the diamine putrescine, the oligoamines spermidine and spermine (Smith and Wyatt, 1981; Smith et al., 1982; Gordonsmith et al., 1983) and the disulphide cystamine (Lewis et al., 1989). The physiological role for this transport system is not know, but it has been suggested that polyamines, which are known to regulate cell growth, may play a role in the differentiation of alveolar type II cells to type I cells (Smith, 1982). It has also been proposed that cystamine represents a source of taurine, which may have an antioxidant role in the lung (Wright et al., 1986; Lewis et al., 1989). Cystamine has also been implicated in playing a role in regulating cellular NADPH levels in response to oxidative stress (Brigelius, 1985). The structural requirements of substrates for this system have been examined and at least two charged nitrogen atoms separated by a distance of at least four methylene groups (about 6.6°A) is essential for uptake (Ross and Krieger, 1981; Gordonsmith et al., 1983; O’Sullivan et al., 1991). It is probable that paraquat, which meets these criteria is recognised as a substrate and thereby accumulated (Smith, 1987). Paraquat accumulation into rat lung slices is reduced in the presence of putrescine in a dose-related manner (Smith and Wyatt, 1981; Karl and Friedman, 1983). Subsequently studies showed that putrescine was accumulated into slices of rat lung by an saturable energy-dependant process with an apparent Km of 7 M and a Vmax of 330 nmol/h/g wet weight of lung. The Km is about 10-fold lower than that

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for paraquat, indicating that the endogenous substrate has a higher affinity for the uptake process than paraquat (Table 83.4). These studies stimulated work to try and identify the specific cell types into which both paraquat and putrescine are accumulated. Slices of rat lung from rats treated with paraquat, which had been shown to cause selective damage to alveolar type I and type II cells, had a decreased ability to accumulate both paraquat and putrescine, suggesting that the transport system resides at least in part in these cell types (Smith et al., 1976; Smith and Wyatt, 1981). This finding is consistent with the autoradiographic studies reported by Waddell and Marlowe, (1980), who showed the distribution of paraquat in mouse lung following iv administration to be consistent with localisation in alveolar type II cells. Studies with rat lung slices in vitro have shown localisation of [3H]paraquat to alveolar type II cells (Wyatt et al., 1988). Similar studies with rat lung slices using [3H]-putrescine, [3H]-spermidine or [3H]-spermine have also shown localisation to alveolar type II cells and in addition provided evidence for accumulation of radiolabel in bronchiolar Clara cells and possibly alveolar type I cells (Wyatt et al., 1988). Similar localisation of [3H]-putrescine was reported by Nemery et al. (1987) and the localisation confirmed by electron microscopy to the type I and type II alveolar epithelial cells and Clara cells (Dinsdale et al., 1991). In contrast, in rabbit lung slices [3H]-putrescine was localised to alveolar type II cells and macrophages but not in Clara cells (Saunders et al., 1988). More recent studies with slices of human lung have established that [3H]putrescine also accumulates into type I and type II alveolar epithelial cells (Hoet et al., 1993). Paraquat accumulation has also been demonstrated in isolated alveolar type II cells from rat and rabbit lung (Forman et al., 1982; Horton et al., 1986, Chen et al., 1992) and in isolated Clara cells from rabbit lung (Horton et al., 1986) suggesting that paraquat transport resides in both cell types. Paraquat is toxic to isolate mouse Clara cells and the addition of putrescine affords some protection (Masek and Richards, 1990). No accumulation of paraquat was however detected in isolated rabbit lung macrophages, although Saunders et al. (1988) have reported putrescine accumulation by rabbit lung macrophages, the basis for this difference is currently not clear but it is now well established that polyamine transport systems are present in a number of transformed and non-transformed blood cells (see Smith et al., 1990). The kinetics of transport of paraquat into isolated type II alveolar epithelial cells has been reported by Chen et al. (1992) using freshly isolated cell suspensions, they found a Km of 88 M with a Vmax of 20 pmol/h/M DNA. They also examined putrescine transport in these alveolar type II suspensions and found a Km of 2.5 M with a Vmax of 33 pmol/h/M DNA. This finding is in broad agreement with that for rat lung slices where the Vmax is very similar for both substrates while the Km for putrescine is higher

Hayes’ Handbook of Pesticide Toxicology

than that for paraquat (Table 83.4). The accumulation of both spermidine and putrescine have been characterised in rat alveolar type II cells in culture (Richards et al., 1987; Kameji et al., 1989; Oreffo et al., 1991). The uptake of spermidine into isolated cells was inhibited by putrescine, spermine and paraquat as described for slices of rat lung. The accumulation of putrescine has also been studied in human alveolar type II cells in culture, the uptake of putrescine and the competitive inhibition by paraquat was essentially the same as that seen in human lung slices (Hoet et al., 1994). Some difficulties have been experienced by several groups in determining the kinetics of transport of paraquat into isolated alveolar type II cells in culture, this may reflect changes to the cell membrane during the isolation procedure, such that the findings in these cells may not accurately reflect that occurring in vivo. A summary of the kinetic constants for the accumulation of both paraquat and putrescine by lung slices and isolate alveolar type II cells for rats and humans is shown in Table 83.4. This data shows that paraquat and putrescine are accumulated by lung slices and alveolar type II cells from both rats and humans and that putrescine has a higher affinity for this system than paraquat. The nature of these transporters is currently not known.

83.2.13  Efflux of Paraquat from the Lung The amount of paraquat that accumulates into the lung is determined by both the rate of accumulation and the rate of efflux from the cells in which it concentrates. The loss of paraquat from rat lung following in vivo administration is slow. There appears to be a rapid phase of elimination over the first 20–30 min following iv administration of paraquat which is then followed by slower loss that obeys first-order kinetics with a half-life of about 50 hr (Sharp et al., 1972). Similar studies by Smith et al. (1978) and Dey et al. (1990) showed a rapid phase of elimination that was similar to that reported by Sharp et al. (1972) while the second phase showed a half-life for paraquat loss from the lung of approximately 20 hr, which was independent of the plasma concentration. Studies in vitro using lung slices from rats dosed in vivo with paraquat also showed a biphasic elimination, with a rapid loss within 30 min presumably reflecting loss from the extracellular space followed by a slower phase with a half-life of 17 hr similar to that seen in vivo (Smith et al., 1981). Thus, the basis for the selective toxicity of paraquat to the lung resides in paraquat’s ability to become concentrated in alveolar type I and II cells and Clara cells. The concentration of paraquat retained in the lung is a combination of that retained during the time of the peak plasma concentration, plus that accumulated via the carrier-mediated process. Paraquat, once accumulated into lung cells, is not then readily lost.

Chapter | 83  Paraquat

Paraquat can be reduced to form a free radical which is stable in aqueous solution in the absence of oxygen (Michaelis and Hill, 1933).

PQ 2  e → PQ.

In the presence of oxygen, in biological systems, the radical will rapidly re-oxidise to the cation with the concomitant production of superoxide anion O2 (Farrington et al., 1973).

PQ.  O2 → PQ 2+  O2.

Thus, once paraquat enters a cell it will undergo alternate reduction followed by re-oxidation a process known as redox cycling. Gage (1968b) first reported that the paraquat cation could be reduced by rat liver NADPH-dependent microsomal flavoprotein reductase to form the radical, with the concomitant oxidation of NADPH. Redox cycling of paraquat has also been reported in microsomal preparations of lung, liver and kidney (Baldwin et al., 1975) and in lung microsomal and slice systems (Adam et al., 1990). Studies using antibodies against NADPH-cytochrome c reductase have shown that paraquat radical formation can be blocked, demonstrating a role for this enzyme in the reduction process (Bus et al., 1974; Horton et al., 1986). Further support for a key role for NADPH-cytochrome c reductase comes from the studies of Kelner and Bagnel1 (1989) using a lymphoblastoid cell line with a specific deficiency in this enzyme which they reported was very resistant to paraquat toxicity. Thus, provided there is sufficient NADPH, as an electron donor, and O2 as an electron acceptor, paraquat will redox cycle inside a cell, generating superoxide anion and consuming NADPH. This reaction is believed to be a key step in the mechanism of paraquat toxicity, however the biochemical consequence of this reaction which leads to lung cell death are complex and still not fully understood. Studies with endothelial cells in culture have indicated that xanthine oxidase can also mediate redox cycling of paraquat to produce superoxide anion (Sakai et al., 1995) indicating that two intracellular enzyme systems are probably involved. Nitric oxide synthase (NOS) also plays a role in paraquat redox cycling as paraquat can use NOS as an electron source to generate O2 at the expense of NO formation (i.e., NOS switches from an oxygenase to a paraquat reductase). Thus paraquat blocks neuronal, endothelial, and macrophage NOS activity and paraquat-induced cytotoxicity in these tissues can be blocked by inhibitors of NOS that prevent NADPH oxidation, but is not attenuated by those that do not (Day et al., 1999). Thus any protective role that NOS may play in protecting lung cells will be compromised by the reduction in NO production produce by paraquat.

Mammalian cells have many enzyme systems which provide them with protection against free radical attack and it is assumed that once these defences have been overwhelmed that cell death occurs. Superoxide dismutase (SOD), is a family of metalloenzymes that can dismutate superoxide anion to hydrogen peroxide and oxygen. O2.  O2. → H 2 O2  O2



The importance of this enzyme in cellular toxicity comes from studies where cellular SOD activity has been genetically modified either by spontaneous mutation or by the transfection of SOD genes. Bilinski and Litwinska (1987) isolated a mutant yeast deficient in SOD activity, which had a greater sensitivity to paraquat than its isogenic wild type. In contrast, Hela cells which possess a higher content of both manganese and copper/zinc SOD had an increased resistance to paraquat (Krall et al., 1988). Transfection of human copper/zinc SOD into various cell lines also lead to resistance to paraquat toxicity (Elroy-Stein et al., 1986; Krall et al., 1988). Recent studies have shown that mice lacking copper/zinc SOD show a marked increase in sensitivity to paraquat (Figure 83.4). Sod/ mice showed a median survival time of about 1.5 days after 10 mg/kg ip, while the Sod/ and Sod/ mice appeared normal at the end of 7 days of observation (Ho et al., 1998). These studies provide strong evidence for a role for superoxide anion radical in the mechanism of cellular toxicity and for the role of copper/Zn SOD in protecting the lungs against paraquat toxicity. However, superoxide anion itself is unlikely to be the ultimate toxic species as it has limited reactivity in biological systems (Halliwell and Gutteridge, 1984). Dismutation of superoxide anion leads to hydrogen peroxide formation which can undergo detoxification by catalase and glutathione peroxidase. Studies with genetically engineered cells have shown that the balance between these two enzymes plays an important role in cellular toxicity of paraquat.

100 80 % survival

83.2.14  Biochemical Mechanisms of Paraquat Toxicity

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+/+ n = 11

60

+/- n = 18

40

-/- n = 14

20 0

0

1

2

3

4

5

6

7

8

Time (days) Figure 83.4  Increased susceptibility of mice lacking Cu/Zn superoxide dismutase to paraquat. The survival times of age-matched, male Sod1/, Sod1/ and Sod1/ mice was determined following ip administration of paraquat at 10 mg/kg . From Ho et al. (1998). Reproduced with permission from Environmental Health Perspectives.

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Increasing intracellular concentrations of SOD to high levels can alter the balance of metabolism of hydrogen peroxide from two electron addition via catalase and glutathione peroxidase to produce water to allow an increase in one electron metabolism to form hydroxyl radical.



H 2 O2 → e → OH. → e → H 2 O ↓ 2e ↓ catalase/GSH peroxidase H2 O

Increasing intracellular SOD content to a very high level, ultimately leads to an increase in toxicity to paraquat in a number of transfected cells or Escherichia coli (Elroy-Stein et al., 1986; Bloch and Ausubel, 1986; Scott et al., 1987; Scott and Eaton, 1996). In contrast, cells having an increase in both SOD and catalase exhibited a greater resistance to paraquat than with just SOD alone (Krall et al., 1988). Generation of hydroxyl radical has been proposed as the critical event in the toxicology of paraquat, this reaction requires the presence of iron and is generated by the Fenton reaction. In this reaction ferrous ions react with hydrogen peroxide to generate hydroxyl radicals.

Fe 2  H 2 O2 → Fe3  OH  OH.

Under physiological conditions free iron predominately exists in the ferric form (Fe3) as a chelate with ADP, ATP and citrate, the reduction of ferric iron may be achieved directly by the paraquat radical (Winterbourn and Sutton, 1984; Sutton et al., 1987) or indirectly by superoxide anion generated from the redox cycling of paraquat (McCord and Day, 1978). A role for transition metals such as iron in the toxicity is supported by studies showing that paraquat toxicity is reduced by removal of iron and enhanced by its addition (Kohen and Chevion, 1985; Sion et al., 1989; Van Der Wal et al., 1990). The role of the iron chelator desferrioxamine in affording some protection against paraquat toxicity will be discussed in the section on antidotes. Many other studies too numerous to mention have been conducted both in vitro and in vivo to explore the effect of altered anti-oxidant status on the toxicology of paraquat. Examples include the role of GSH and GSH reductase (Bus et al., 1976a; Keeling, et al., 1982; Hardwick et al., 1990), the role of selenium deficiency, vitamin E and glutathione peroxidase (Bus et al., 1976b; Cagen and Gibson, 1977; Block, 1979; Omaye et al., 1978; Kelner et al., 1995) and the role of metallothionein (Satoh et al., 1992; Lazo et al., 1995). Metallothionein appears to have play a role as a free radical scavenger in addition to its well established role as a heavy metal chelator, metallothionein has been reported to quench both superoxide anion and hydroxyl radicals,

with a significantly higher reactivity towards hydroxyl radicals (Thornalley and Vasak, 1985). Genetically engineered animals have been used as tools to try and elucidate the function of the various antioxidant defence mechanism against paraquat-induced oxidant injury. In addition to the discussion above regarding mice deficient in copper/zinc SOD, Sato et al., (1996) found mice deficient in metallothionein I and II genes to be more susceptible to paraquat toxicity. Glutathione peroxidase deficient mice show a increased susceptibility to paraquat toxicity with a mean survival time of 5 hr compared to the wild type of 69 hr following an ip dose of 50 mg/kg (Cheng et al., 1998). Mice over expressing glutathione peroxidase are more tolerant to paraquat toxicity, wild type mice given a 125 mg/kg ip dose of paraquat died within 5 h while the mice over expressing the enzyme lived until about 54 hr (Cheng et al., 1998). Peroxiredoxin 6 (Prdx6), a bifunctional 25-kDa protein with both GSH peroxidase and phospholipase A2 activities also plays an important role as an antioxidant and is expressed in all major organs, with a particularly high levels in the lung. Studies in Prdx6 null mice has shown they are more sensitive to the effects of hyperoxia or paraquat. It has been postulated that Prdx6 functions in antioxidant defense mainly by facilitating repair of damaged cell membranes via reduction of peroxidized phospholipids (Manevich and Fisher, 2005). As discussed earlier paraquat causes both acute lung injury and in some animals goes on to cause fibrosis. Recent studies using gene expression profiling have looked at the response of the lung to paraquat during the early phase within the first 24 h following exposure and during the destructive phase and the onset of fibrosis. Genes associated with oxidative stress and redox cycling eg. haem oxygenase-1, NAD(P)H dehydrogenase, quinone1 (NQO-1); Prdx6; thioredoxin reductase 1; glutamate-cysteine ligase, transferrin receptor were markedly up regulated 2–8 h after a median lethal dose of 20 mg/kg paraquat to rats (Tomita et al., 2006a; Mainwaring et al., 2006). A similar stress response was seen in mice 6 h after nasal instillation of paraquat into the lung (Tomita et al., 2007). Interestingly at this early time genes associated with pathways involved in fibrosis were also upregulated such as TGF-2 and its receptor, SMAD6, PDGF receptor, TIMP, immunophilin and integrin  (Mainwaring et al., 2006). In a multiple dosing model in rats or mice, gene changes associated with fibrogenic growth factors such as TGF, PDGF-A, acidic fibroblast growth factor, procollagen were detected in the lung 5 days to3 weeks after exposure (Satomi et al., 2004, 2007; Ishida et al., 2006; Tomita et al., 2007). Satomi et al. (2007) also examined gene changes in the lungs of paraquat-treated rats that developed fibrosis with those rats that did not develop fibrosis and identified 36 genes that were significantly changed between the two groups and in particular noted the pro-apoptotic gene Bad and suggested that this gene could play a key role in determine which rats develop fibrosis.

Chapter | 83  Paraquat

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Figure 83.5 shows a schematic representation of the key requirements to enable paraquat to enter a cell and the subsequently redox cycling steps believed to lead to cytotoxicity. Three hypotheses have been proposed to account for the ensuing cytotoxicity, one involving lipid peroxidation, another the oxidation of NADPH and the third mitochondrial toxicity, none of these hypotheses are mutually exclusive. The current weight of evidence supports the redox cycling of paraquat in both the cytosol and mitochondria leading to oxidative stress, resulting in an alteration in the redox state of the cell, a reduction in energy production and lipid peroxidation.

1983) and in vivo (Bus et al., 1976b; Reddy et al., 1977; Burk et al., 1980). However, others have questioned its significance in the toxicity, for example Steffen et al., (1980) only found a small increase in the exhalation of ethane (a marker of lipid peroxidation) in rats suffering from respiratory distress following exposure to paraquat and oxygen. Similarly, others have been unable to find evidence of lipid peroxidation in the lungs of mice given large doses of paraquat (Shu et al., 1979; Younes et al., 1985) or it is only detected as a late event in the toxicity (Ogata and Manabe, 1990). So the question remains as to whether lipid peroxidation is a cause, or a consequence, of the toxicity. These contrasting findings in vivo may also reflect the difficulty in detecting a small but critical increase in lipid peroxidation in the alveolar type I and II cells and Clara cell that are only a small population of the total cells in the lung. 4-Hydroxy-2-nonenal (4-HNE), is a lipid peroxidation marker and use of an antibody to detect 4-HNE protein adducts has shown an increase in immunoreactivity in the lung and kidney but not liver of C57bl/6 mice given a MLD of paraquat (Kurisaki and Hiraiwa, 2009). Studies in the brain of C57Bl/6 mice exposed to three weekly injections of paraquat, showed that the vast majority of midbrain 4-HNE-immunoreactive cells were dopaminergic (tyrosine hydroxylase-immunoreactive), suggesting a relationship between lipid peroxidation and neuronal death

83.2.15  Lipid Peroxidation Hypothesis Bus and co-workers (1974; 1976a) proposed the sequential generation of superoxide anion and hydroxyl radical and the initiation of lipid peroxidation as the mechanism of cellular toxicity of paraquat. However, there is little direct evidence which demonstrates lipid peroxidation occurs in the lung of animals dosed with paraquat before there is morphological evidence of cell damage. Paraquat-induced lipid peroxidation has been demonstrated in vitro in broken cell systems and isolated cells from the lung and liver (Bus et al., 1976a; Kornbrust and Mavis, 1980; Trush et al., 1981; Sata et al., 1983; Saito et al., 1985; Sandy et al., 1986; Aldrich et al.,

Hexose monophosphate shunt

PQ

0.702 nm

–ve

1

–

+

NH3—(CH2)4—NH3

Alveolar epithelial cell membrane

O2 + +

– O•2 – O•2

2H+

H2O2 .......................... 1

Fe2++ O2 .................. 2 OH•+OH–+Fe3+...... 3 4

Glutathione peroxidase

GSSG

– O•2 Fe3+

Fe2+ + H2O2

5 H2O

PQ+•

O•2 then

0.622 nm

3

2

Putrescine +

NADP+

NADPH PQ2+

r

–ve >0.5 nm

epto

+ NCH3

Rec

+ CH3N

GSH

Glutathione reductase

Lipid peroxidation Cell death

NADPH

NADP+

Hexose monophosphate shunt

Figure 83.5  Mechanism of toxicity of paraquat. A schematic representation of the mechanism of toxicity of paraquat. 1structure of paraquat and putrescine showing the geometric standards of the distance between the n atoms; 2transport system which recognises paraquat, minimum separation of charge of approximately 0.5 nm; 3redox cycling of paraquat utilising NADPH; 4formation of hydroxyl radical leading to lipid peroxidation; 5detoxification of H2O2 via glutathione reductase/peroxidase couple, utilising NADPH. From Smith (1987). Reproduced with permission from Human Toxicology.

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(McCormack et al., 2005; 2006). These later studies demonstrate the presence of oxidative stress (lipid peroxidation) in the nigrostriatal region of the mouse brain following paraquat administration.

83.2.16  Oxidation of NADPH Hypothesis Intracellular redox cycling of paraquat results in the oxidation of NADPH leading to cellular depletion such that those cells that selectively accumulate paraquat can longer function normally. Fisher et al., (1975) first suggested that the redox potential of lung cell may be altered by the redox cycling of paraquat. A marked stimulation of the activity of the pentose phosphate pathway in the lung has been observed following exposure to paraquat (Fisher et al., 1975; Rose et al., 1976b; Bassett and Fisher, 1978; Keeling et al., 1982 and Fisher and Reicherter, 1984). Since this pathway represents the major cellular source of NADPH, it is inferred that this response represents an attempt by lung cells to maintain their levels of reducing equivalents under conditions of oxidative stress. In those cells in which paraquat is accumulated, the concentration may be very high and result in very fast rates of NADPH oxidation. If the rate of consumption exceeds the rate of formation via the pentose phosphate pathway, the concentration of NADPH will fall below that required to maintain cell viability. Witschi et al. (1977) first demonstrated that the NADPH/NADP ratio in the lungs of rats dosed iv with paraquat was decreased suggesting that oxidation of the reduced nucleotide had occurred. Later studies by Keeling and Smith, (1982) demonstrated that the shift in NADPH/NAD ratio in the lung following sc administration of paraquat was the result of NADPH loss from the lung. A consequence of depletion of cellular NADPH is that the cell shuts down it synthetic pathways which are dependent on this nucleotide, such as the synthesis of fatty acids (Keeling et al., 1982). A loss of NADPH may also have particular importance for alveolar type II cells which produce pulmonary surfactant (Brigelius et al., 1986). NADPH is also consumed in an attempt by the lung to detoxify hydrogen peroxide that is formed via the glutathione peroxidase/reductase enzyme system (Figure 83.5) to regenerate reduced glutathione (GSH) from its oxidised form (GSSG). In general large changes in lung GSH and GSSG are not seen after paraquat administration (Bus et al., 1976a; Reddy et al., 1977; Shu et al., 1979; Keeling and Smith, 1982) this may explain why lipid peroxidation has not been conclusively demonstrated in vivo as this would not become apparent until both NADPH and GSH were markedly reduced. However, formation of protein mixed disulphides are increased in the lung in vivo (Keeling et al., 1982; Keeling and Smith, 1982) and in perfused liver (Brigelius et al., 1982). These changes in protein mixed disulphides in the lung are presumably a response to oxidative stress and may not be critical to the cellular toxicity. This notion

Hayes’ Handbook of Pesticide Toxicology

is supported by studies with the bipyridyl diquat which can undergo redox cycling in the lung (Rose et al., 1976b ; Witschi et al., 1977). Diquat also produced increases in protein mixed disulphide content in the lung without affecting NADPH content at a dose that did not cause lung injury (Keeling and Smith, 1982). This indicates that NADPH depletion subsequent to redox cycling is a critical step in the mechanism of paraquat toxicity.

83.2.17  The Role of Mitochondria   in the Toxicity Another hypothesis that has been proposed is that paraquat toxicity is due to mitochondrial damage, based on morphological findings of early mitochondrial changes in alveolar type II cells (Hirai et al., 1985) Ultrastructural studies of the time course of development of paraquat-induced lung injury have also reported early changes to mitochondria such as swelling and altered staining density (Smith and Heath, 1974a; Sykes et al., 1977; Keeling et al., 1981). These mitochondrial changes were also observed in the lungs of rats exposed to paraquat and 85% oxygen, which enhances paraquat toxicity to the lung (Keeling et al., 1981). However, as discussed with regard to the lipid peroxidation hypothesis the question is are the effects on mitochondria a cause, or a consequence, of paraquat toxicity. Early studies with isolated liver mitochondria reported only minor changes in mitochondria respiration by paraquat (Gage, 1968b). More recent studies have reported that paraquat cation can be reduced by NADH-ubiquinone oxidoreductase (Complex I) located on the inner mitochondrial membrane (Fukushima et al.,1993; Shimada et al., 1998), these authors also showed that paraquat was able to stimulate lipid peroxidation in submitochondrial particles (Yamada and Fukushima, 1993). These findings show that mitochondria have the potential to generate superoxide anion from paraquat provided it can gain access. In general, studies with intact mitochondria support the original findings of Gage, (1968b) showing that little or no effects are seen (Lambert and Bondy 1989, Constantini et al., 1995) unless very high concentrations of paraquat are present (Kopaczyk-Locke, 1977;Yamamoto et al., 1987; Thakar and Hassan, 1988; Palmeira et al., 1995). Paraquat has been shown to induce a Ca2-dependent permeability transition of the inner mitochondrial membrane leading to membrane depolarisation, uncoupling and matrix swelling in isolate rat liver mitochondria (Constantini et al., 1995). This opening of the membrane permeability pore does not occur in the absence of added Ca2 and requires the presence of rotenone, leading one to question the relevance of this observation to the in vivo situation. Based on these studies it seem likely that any intracellular increases in Ca2 would be only occur once paraquat had entered the lung cell, undergone redox cycling and altered mixed disulphide status. In summary, mitochondrial damage has been observed in the lung

Chapter | 83  Paraquat

prior to cell death, it seems likely that this response may be secondary to changes taking place in the cytosol. Recent studies in rat brain mitochondria have shown that mitochondria are the principal cellular site of paraquatinduced H2O2 production, and that it requires the presence of a respiratory substrate, the involvement of complex III of the electron transport chain and depends on the mitochondrial inner transmembrane potential (Castello et al., 2007). These authors also showed that paraquat can cross the mitochondrial membrane and that this uptake is blocked by inhibitors of the respiratory chain such as antimycin A and uncouplers. The effect of paraquat on mitochondrial H2O2 production was also confirmed in rat mid-brain cultures of neurons. Subsequent studies in mitochondria from rat liver and heart and from yeast have confirmed that paraquat can cross the inner mitochondrial membrane and cause H2O2 production, however in contrast to Castello et al., (2007) their findings support a role for complex I not III as the major site of paraquat reduction (Cocheme and Murphy, 2008). These findings indicate that if paraquat enters a cell then it can also cross the mitochondrial membrane and thereby induce oxidative stress.

83.2.18  The Involvement of Oxygen As discussed earlier, the redox cycling of paraquat to form superoxide anion requires oxygen and hence oxygen plays a critical role in the toxic process. It has been known for many years that hyperoxia is toxic to the lung, causing damage to endothelial cells through a mechanism that involves the formation of reactive oxygen species (Frank and Massaro, 1979; Jenkinson, 1982). One of the therapeutic measures for anoxia in human cases of paraquat poisoning was the addition of air supplemented with oxygen (see treatment of human poisoning). However, it has been shown that increasing the oxygen concentration potentiates the lethality of paraquat to rats (Fisher et al., 1973; Douze and van Heijst, 1977; Kehrer et al., 1979; Keeling et al., 1981) increasing the injury to the lung. While the converse is true, rats exposed to paraquat in a hypoxic environment are protected relative to those exposed to paraquat in air (Rhodes et al., 1976). Detailed histopathology on the lungs of rats exposed to paraquat alone or paraquat in an atmosphere of 85% oxygen, showed that the damage was primarily localised to the alveolar type I and II cells with little evidence of endothelial cell damage, showing that oxygen potentiated paraquat toxicity (Keeling et al., 1981). These findings have recently been reproduced using isolated rat and human alveolar type II cells exposed to either paraquat in air or paraquat and increasing concentrations of oxygen. Increasing the oxygen concentration in the atmosphere potentiated the toxicity of paraquat, while lowering the oxygen concentration to 10% afforded some protection (Hoet et al., 1997). The mechanism underlying this synergistic effect of oxygen on paraquat toxicity is not entirely clear, it seems unlikely

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that oxygen would normally be rate limiting for paraquat to redox cycle, a more likely explanation is that the cellular defence mechanisms that protect against oxygen and paraquat toxicity are more rapidly overwhelmed. In summary, the key events leading to cellular toxicity are (1) accumulation of paraquat into the cell and (2) its ability to redox cycle and produce oxidative stress in both the cytosol and mitochondria. It seems likely that a combination of mitochondrial dysfunction, depletion of NADPH plus the generation of hydroxyl radical leading to lipid peroxidation and is involved but the precise temporal relationships have not as yet been established.

83.2.19  Effects on the Kidney The major route of elimination for paraquat once it has entered the blood stream is via the kidneys where it is actively secreted by organic cation transport systems (see review by Chan et al., 1998). This process becomes saturated at fairly low plasma concentrations (3–4 nmol/ml; 0.5–0.75 g/ml) in the rat (Chan et al., 1997). At higher plasma concentrations paraquat is nephrotoxic, large oral or systemic doses administered to rats or mice producing morphological changes to the proximal renal tubules, including hydropic degeneration with occasional evidence of necrotic epithelial cells and of renal tubular regeneration (Clark et al., 1966; Lock and Ishmael, 1979). Chronic exposure to mice via their drinking water showed ultrastructural evidence for proliferation of smooth endoplasmic reticulum and the presence of lipid containing bodies in proximal tubule cells (Fowler and Brooks, 1971). Renal tubular necrosis is more marked in the dog and rabbit following large toxic doses with clear evidence of degeneration of proximal tubular cells with the presence of casts in the tubular lumen (Clark et al., 1966; McElligott, 1972; Yonemitsu, 1986; Giri et al., 1982; Nagata et al., 1992a). Prior to the onset of renal tubular necrosis, paraquat-induced renal functional changes including diuresis, albuminuria, glucosuria and elevations in plasma urea and creatinine in the rat (Lock and Ishmael, 1979), dog (Giri et al., 1982; Nagata et al., 1992a) and cynomologus monkey (Purser and Rose, 1979). The precise mechanism of renal functional impairment is not known, it probably involves altered renal haemodynamics as well as accumulation of paraquat into proximal renal tubules leading to cellular necrosis. There is some evidence that paraquat may reduce renal blood flow based on the finding of elevated renal plasma renin activity in the dog after dosing (Giri et al., 1982) and hypovolaemia in the rat (Lock, 1979). Paraquat enters renal tubular cells by an organic cation transport system (OCT2, see 70.2.11), thereby enabling it to concentrate to many time that present in the plasma (Figure 83.3)(Ecker et al., 1975a; Lock and Ishmael, 1979; Hawksworth et al., 1981; Wright and Wunz, 1995; Groves et al., 1995; Chan et al., 1996a, 1996b, 1997; Chen et al., 2007). Exit via the apical or brush border membrane

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is via the multidrug and toxic extrusion transporter MATE-1 which is a sodium/H exchange (Chen et al., 2007) (Figure 83.3). The accumulation can be blocked by other organic cations such as tetraethylammonium and quinine but is not affected by the polyamines, putrescine or spermine (Groves et al., 1995; Chan et al., 1996a). Thus the accumulation of paraquat into renal tubular cells occurs via a different transport system to that which leads to its accumulation in the lung. Once inside a renal tubular cell paraquat can redox cycle (Baldwin et al., 1975, Tomita, 1991) producing superoxide anion and hence trigger the cascade of biochemical events leading to cytotoxicity similar to that discussed for the lung (Lock and Ishmael, 1979; Molck and Friis, 1997). Further support for redox-cycling causing damage to the kidney comes from gene expression studies where a median lethal dose of paraquat given to rats lead to an upregulation of haem oxygenase 1, and metallothionein-1 in the kidney 3 h after dosing (Tomita et al., 2006b). These findings are similar to those reported for gene expression changes in the lung of rats exposed to paraquat, see 70.2.14. Regardless of the mechanism, the consequence of a reduced renal excretion is that more paraquat is available in the plasma, to accumulate into the lung. Thus, maintenance of renal function to facilitate paraquat excretion from the body is critically important for cases of human poisoning (see later).

83.2.20  Effects on the Central   Nervous System No clinical signs of neurotoxicity or neuropathological changes have been reported following oral gavage or dietary administration of paraquat to rodents or dogs in the studies required for registration of paraquat (IPCS, 1984; JMPR, 2003). Paraquat as a di-cation does not readily cross the blood-brain barrier and enter the rat brain after either oral or systemic administration (Rose et al., 1976a; Dey et al., 1990; Corasaniti et al., 1991; Corasaniti and Nistico, 1993; Naylor et al., 1995; Widdowson et al., 1996a and b). The concentration associated with the rat brain is always lower than that in the plasma and decreases with time, the initial concentration detected in the brain may be largely associated with blood (Rose et al., 1976a; Dey et al., 1990; Naylor et al., 1995). Paraquat was however detected in brain regions such as the olfactory bulb, area postrema and hypothalamus, which do not possess an effective bloodbrain barrier. Autoradiographic studies have detected paraquat in these regions and in the cerebrospinal fluid (ventricles and choroid plexus) but the concentrations were low and only represent a very small percentage of the administered dose, about 0.05% at the time of maximal blood concentration 1 hr after dosing (Waddell and

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Marlowe 1980; Naylor et al., 1995). Immunohistochemical localisation of paraquat in rat brain has shown it is present in capillary walls and glial cells but was not detected in neurones (Nagao et al., 1991; McCormack and DiMonte, 2003). More recent studies have examined the entry of paraquat into the brain of C57Bl/6 mice, following ip administration of 10 mg/kg 2 or 3 times a week, brain levels plateaued after 18 doses by which time there was some morbidity and mortality. Between 18 and 36 doses striatal levels of dopamine and tyrosine hydroxylase were decreased. Administration of paraquat in the drinking water at 0.03–0.05 mg/ml did not result in any mortality but paraquat again showed a time and concentration dependent increase in the brain (Prasad et al., 2009). Studies in the rat, using parenteral doses of paraquat at or above the MLD (20–100 mg/kg, ip) produced signs of neurotoxicity with muscle fasciculation, some tremors and “wet-dog” shakes and at the higher doses myoclonus, typically within 30 min of dosing (Bagetta et al., 1992; Corasaniti et al. 1992; Hara et al., 1993), which is the time of peak blood and brain concentrations. These authors also reported neuronal cell necrosis in the pyriform cortex of these animals 24 h after dosing (Bagetta et al., 1992; Corasaniti et al., 1992). The neuronal cell necrosis could be reduced by administration of atropine but not methylatropine (Bagetta et al., 1992), suggesting some involvement of central muscarinic receptors. No effects were seen after 5 mg/kg ip paraquat. The basis for the selective injury to the pyriform cortex is currently not known, but it does not reflect the brain region with the highest concentration of paraquat (Corasaniti and Nistico, 1993; Naylor et al., 1995). Others have reported that paraquat (20 mg/kg, sc) does not produce neuronal cell necrosis in the pyriform cortex of perfused-fixed material from rats 24 and 48 h after dosing (Naylor et al., 1995; Widdowson et al., 1996a) and have suggested the effect reported by the Italian group may be a fixation artefact. The precise basis for this variance is currently not understood. Similarly, daily oral dosing of paraquat at 5 mg/kg/day for 14 days to rats produced no evidence of neuronal cell necrosis, despite particular emphasis on the pathology of the pyriform cortex, nigrostratial region and hypothalamus, or behavioural changes indicative of neurotoxicity (Widdowson et al., 1996b). Direct administration of paraquat into the ventricles or infusion into certain brain regions produced signs of neurotoxicity in rats which were associated with neuronal cell damage (De Gori et al., 1988; Calo et al., 1990; Bagetta et al., 1992, 1994; Corasaniti et al., 1992; Yoshimura et al., 1993; Liu et al., 1995; Liou et al., 1996). These effects were seen at low doses of paraquat 2–20 g injected. These observations lend support to the view that little paraquat enters the brain following systemic administration (20 mg/kg, sc or 4,000 g /200 g rat) or oral administration (126 mg/kg or 25,200 g/200 g rat) as no neuronal cell toxicity was seen at these doses.

Chapter | 83  Paraquat

Comparisons have been drawn to the structural similarity between paraquat and 1-methyl-4-phenylpyridinium ion (MPP), the active metabolite of 1-methyl-4-phenyl1,2,3,6-tetrahydropyridine (MPTP) which can induce a Parkinson-like syndrome in monkeys and humans. Administration of MPTP to susceptible animal species produces selective damage to dopaminergic neurons in the substantia nigra leading to a marked loss of dopamine and clear signs of neurotoxicity. The mechanism for MPTP toxicity (see Markey et al., 1986; Tipton and Singer, 1993) is due to its ability to cross the blood-brain barrier and enter glial cells where it can undergoes oxidative metabolism by the enzyme monoamine oxidase B to form MPP. This metabolite then accumulates selectively into dopaminergic neurons via the dopamine transport system, leading to inhibition of mitochondrial respiration which ultimately leads to the demise of the neurone. Structure-activity relationships suggest that, despite their apparent similarity, paraquat and MPTP are two very different chemicals (Koller, 1986; Miller, 2007): MPTP is uncharged and lipophilic and thereby able to cross the blood-brain barrier, whereas paraquat is charged and hydrophilic and does not readily enter the brain. Also, MPTP is a monoamine whose metabolite MPP, the proximate toxin, is able to use a specific uptake system, particularly in the substantia nigra, whereas paraquat is a diamine. It is also very relevant that administration of MPP to experimental animals did not produce neurotoxicity, due to its poor entry across the blood-brain barrier (Tipton and Singer, 1993). Thus like paraquat, MPP does not readily enter the brain. Consistent with this systemic administration of paraquat to C57 black mice or rats did not lead to dopamine depletion or neuronal cell death in the striatum, like that seen with MPTP (Perry et al., 1986; Widdowson et al., 1996b). Others have reported changes in brain dopamine content following paraquat administration to mice (Endo et al., 1988; Fredriksson et al., 1993). In the latter case, paraquat was administered to pups on days 10 and 11 after birth at a time when the brain is undergoing rapid growth and hence might be a more vulnerable to chemical insult. The authors reported a small loss of dopamine and its metabolites and a decreased behavioural activity when measure at about 4 months of age (Fredriksson et al., 1993). This suggests that the developing brain is potentially more sensitive to insult, however adverse effects have not been detected in developmental toxicity or multigeneration studies, where paraquat was given to pregnant rats and their offspring (see earlier section). Attempts to try and reproduce the findings of Fredriksson et al. (1993) in C57 black mice, in another laboratory, have not proved possible (David Ray, personal communication). In the last decade animal studies have focussed on the entry of paraquat into the brain, its potential to cause neuronal cell death in the substantia nigra and its interaction with the fungicide maneb. The main reason for this is the putative link between pesticides and Parkinson’s disease

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(Brown et al., 2006; Dreschel and Patel, 2008; Miller et al., 2009). Initial studies in adult C57Bl/6 mice administered paraquat 10 or 30 mg/kg, once a week for 3 weeks reported behavioural changes and loss of dopaminergic neurons in the substantia nigra of the pars recta (Brooks et al., 1999; McCormack et al., 2002). This was followed by combination studies with pesticides, where adult C57Bl/6 mice were treated twice a week ip, for 6 weeks with 10 mg/kg paraquat and 30 mg/kg maneb. This regimen showed a more marked effect on motor activity, striatal dopamine and dihydroxyphenylacetic acid content and tyrosine hydroxylase activity than when either compound was given separately (Thiruchelvam et al., 2000). This group then reported in collaboration with DiMonte’s group that paraquat alone given once a week for three weeks at 10 mg/kg caused neuronal cell death in the substantia nigra pars compacta, but not in the substantia nigra pars reticularta or in the hippocampus, of 6 week, 5 and 18 month old C57Bl/6 mice using stereological counting of tyrosine hydroxylase stained neurons (McCormack et al., 2002). However no significant depletion in striatal dopamine was observed showing a discrepancy between the pathology and neurochemistry in these dopaminergic neurones. The finding of neuronal cell loss in the substantia nigra of mice dosed with paraquat using stereology has subsequently been confirmed in another laboratory (Fernagut et al., 2007). One dose of paraquat is not sufficient to produce neuronal cell loss in the substantia nigra of mice, however Purisai and coworkers (2007) showed it did cause microglial cell activation, and suggested that this might prime the pre-disposed neurons to subsequent injury. To support their case they showed that inhibiting the microglial response with an anti-inflammatory drug protected mice against subsequent doses of paraquat with regard to neuronal cell loss. In contrast, when microglial cells were activated by pre-treatment with lipopolysaccharide a single dose of paraquat led to dopaminergic cell loss (Purisai et al., 2007). This group also showed that mice lacking functional NADPH oxidase were resistant to paraquat-induced dopaminergic cell loss. Overall these findings indicate that sufficient paraquat can enter the brain and hence the substantia nigra and cause selective dopaminergic cell loss in C57Bl/6 mice following multiple weekly doses of 10 mg/kg ip, although the neuronal cell loss is never more than 25–30% (McCormack et al., 2005), suggesting a sensitive pool of dopaminergic neurons. However, daily intranasal administration of paraquat to C57Bl/6 mice for 30 days caused no effect to the nigrostriatal system while administration of MPTP by the same route caused extensive loss of tyrosine hydroxylase positive cells in the substantia nigra and striatum (Rojo et al., 2007). In 2001 a Japanese group reported detecting paraquat in the dialysate of rats undergoing brain microdialysis following subcutaneous administration of paraquat. In contrast MPP did not penetrate the blood-brain barrier in either

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control or paraquat-treated rats (Shimizu et al., 2001). These data suggested that paraquat may be entering via a carrier mediated process and the authors went on to show that the neutral amino acid L-valine could reduced the blood-brain barrier penetration of paraquat while putrescine, a good inhibitor of paraquat uptake into the lung, was without effect (Shimizu et al., 2001). This finding is very suprising, that a di-cation can enter the brain via a neutral amino acid transporter and raises questions about the specificity of these transporters. It appears that paraquat can enter lung cells by a diamine transporter, renal cells via an organic cation transporter and cross the blood-brain barrier via the L-amino acid transporter. Could the charge on paraquat in some way be masked to allow entry into the brain? This finding was confirmed and extended by McCormack and DiMonte, (2003) who using an antibody to detect paraquat showed that its entry in the brain of mice was reduced by L-valine, L-phenylalanine and the antiparkinsonian drug L-dopa but not by D-valine or L-lysine. Pretreatment with chemicals that use the neutral amino acid transport system also protected against paraquat-induced loss of nigrostriatal dopaminergic cells. Thus another laboratory has confirmed the findings reported by Shimizu and coworkers that pretreament of either rats or mice with these amino acids can reduce paraquat entry into the brain. Intraneuronal aggregation of -synuclein is a known feature of Parkinson’s disease and studies in vitro have shown that paraquat can accelerate the formation of -synuclein fibrils in a dose-dependent manner. Mice exposed to three weekly injections of paraquat also showed a consistent increase in -synuclein in the brain when measured 2 days after each weekly injection (Manning-Bog et al., 2002). The relevance of -synuclein accumulation to paraquat-induced nigrostriatal cell loss has been examined in control mice and mice overexpressing -synuclein and it showed that mice overexpressing -synuclein were completely protected against the neuronal cell loss (Manning-Bog et al., 2003). Thus as far as upregulation of -synuclein is concerned, it may not be a toxic mediator of paraquat-induced dopaminergic cell death, although a note of caution is necessary as increased expression of -synuclein is toxic in its own right to dopaminergic neurons in humans (Singleton et al., 2003). In contrast, Fernagut et al. (2007) reported that overexpression of -synuclein in mice did not alter the effect of paraquat on dopaminergic neurons in the substantia nigra pars compacta, thus the role of -synuclein is unclear. Once paraquat has entered the brain, why is it selective for dopaminergic neurons in the nigrostriatal region? One hypothesis is that these neurons may be more susceptible to oxidative stress (McCormack et al., 2006). However there is no evidence todate showing that paraquat is present in the striatum at a higher concentration than in other brain regions (Naylor et al., 1995; Barlow et al., 2003; Prasad et al., 2009), although these neurons may have a reduced protection against oxidative stress. With MPP this is due to its

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ability to enter these neurons via the dopamine transporter where it can concentrate and inhibit complex 1 of the mitochondrial respiratory chain. However, paraquat is neither a substrate nor inhibitor of the dopamine transport system and does not act by inhibiting mitochondria complex 1 (Richardson et al., 2005; Miller, 2007). However, this later finding has been challenged by studies showing that complex I is the predominant site in mitochondria for paraquatinduced reactive oxygen species (Cocheme and Murphy, 2008) suggesting MPP, rotenone and paraquat do have a common site of action in mitochondria. In vitro studies in neuronal cell lines have shown that following paraquat-induced oxidative stress activation of Jun N-terminal MAPK-mediated caspase-3 dependent cell death occurs (McCarthy et al., 2004; McCormack et al., 2002; Peng et al., 2004; Ramachandiran et al., 2007) and further that this does not occur with MPP, suggesting a different signalling cascade for paraquat. All of the above studies have been conducted in C57Bl/6 mice raising the question of the response in other species. Studies have shown that paraquat can lead to nigrostriatal dopaminergic cell loss in the rat, to a similar extent to that seen in the mouse, although much greater exposure is required with doses of paraquat of 10 mg/kg, ip, twice a week for 4–24 weeks (Ossowska et al., 2005) This group also report an increase in turnover of dopamine without any depeletion in the total content, indicating that the remaining nigrostriatal cells can compensate for the cell loss (Ossowska et al., 2006; Kuter et al., 2007). Combination studies of paraquat (10 mg/kg) and maneb (30 mg/kg) twice a week for 4 weeks to adult rats, resulted in loss of dopaminergic neurons and microglial cell activation in the substantia nigra (Cicchetti et al., 2005). Similar studies in 6 month old rats showed the same response although they were more sensitive half of the rats developed fatal lung injury (SaintPierre et al., 2006). The entry of paraquat into rat brain can be prevented or reduced as discussed earlier by administration of the amino acid L-valine (Shimizu et al., 2001; Chanyachukul et al., 2004). Nicotine has been shown to partially protect mice against paraquat-induced nigrostriatal damage indicating an involvement of 62 nACh receptors, without affecting dopaminergic activity (Khwaja et al., 2007). Further studies by this group with paraquat in monkeys showed a somewhat different response, paraquat down regulating striatal nACh receptor-mediated dopaminergic activity (O’Leary et al., 2008). Thus it is important to understand the response of higher species to paraquat when trying to determine the likely response of humans. In this regard it has been recently reported that paraquat is excluded from entry into the brain of rhesus macaque monkeys using in vivo positron emission tomography imaging (Bartlett et al., 2009). These authors assessed the uptake of [11C]- para­ quat in adult monkeys and found that brain concentrations including those in dopaminergic areas were consistent

Chapter | 83  Paraquat

with the blood volume in those structures. This raises the question of whether monkeys are susceptible to paraquatinduced nigrostriatal dopaminergic cell loss. Todate, there have been no reports showing neuronal cell loss in higher mammals, raising the pivotal question of the relevance of these findings to humans. In summary, over the last 10 years it has become clear that paraquat can enter the brain of C57Bl/6 mice following multiple parenteral administrations at doses of 10 mg/kg given once a week and cause neuronal cell loss in the substantia nigra. However, administration by nasal instillation, a more relevant route for potential human exposure, showed no injury to the substantia nigra. Rats also develop neuronal cell loss in the substantia nigra after paraquat exposure, although much greater exposure is required than for the C57Bl/6 mouse, while the position with monkeys is still unclear.

83.2.21  Effects on Other Organs Following oral ingestion of paraquat by humans, ulceration of the pharyngeal, oesophageal and gastric mucosa has been reported (see section on human toxicology). In animal studies there is often no direct contact with these tissues when gavage dosing is employed. However, focal necrosis of the gastrointestinal tract has been observed in primates, demonstrating the topical irritant nature of high oral doses of paraquat (Murray and Gibson, 1972). Paraquat administration to the rat produced an increased synthesis of liver glycogen and an increase in blood glucose that appeared to be mediated by the adrenal, since adrenalectomy prevented these changes (Rose et al., 1974b). These effects seen following paraquat and the related bipyridyl diquat are thought to be due to catecholamine release and high circulating concentrations of corticosteroids (Rose et al., 1974b). This response is thought to be unrelated to the pulmonary damage produced by paraquat, but may account for some of the effects seen with paraquat on the adrenal and lymphoid tissues such as the spleen and thymus (Clark et al., 1966; Butler and Kleinerman, 1971; Fisher et al., 1973). The increase in circulating corticosterone seen with paraquat can also be prevented by lesioning the area postrema (Edmonds and Edwards, 1996). This area of the brain also controls the taste aversion to paraquat seen in rats (Dey et al., 1987; Edmonds and Edwards, 1996). Liver damage is not a major finding after paraquat administration, after large doses some central lobular necrosis has been reported in most species examined (Clark et al., 1966; Murray and Gibson, 1972; Cagen et al., 1976; Giri et al., 1982; Nagata et al., 1992a). Since paraquat is delivered to the liver following dosing and hepatocytes possess the relevant enzymes to facilitate redox cycling, presumably paraquat does not normally accumulate in hepatocytes to a sufficient concentration to overwhelm the protective a­ntioxidant defence enzymes and produce necrosis.

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However, both mice and rats made selenium deficient show marked liver injury following paraquat administration (Cagen and Gibson, 1977; Burk et al., 1980) supporting the view that selenium dependent enzymes such as glutathione peroxidase play an important protective role. These findings are consistent with recent studies using transgenic mice where glutathione peroxidase has either been deleted or over expressed showing that this selenium dependent enzyme plays a key role in paraquat-induced tissue injury (Cheng et al., 1998; De Haan et al., 1998).

83.2.22  Treatment of Poisoning in Animals Over the past 30 years a variety of attempts to modify the toxicity of paraquat in experimental animals have been examined. To date the only approach that has been shown to clearly reduce mortality in rats is purgation of the gastrointestinal tract with a diatomaceous clay (bentonite or Fuller’s earth) along with a cathartic (e.g. magnesium sulphate)(Clark, 1971; Smith et al., 1974). Attempts to modify paraquat toxicity have been based on its known mechanism of toxicity and will be briefly discussed under the following headings: (a) prevention of absorption from the gastrointestinal tract; (b) removal from the blood stream; (c) prevention of accumulation into the lung; (d) attempts to scavenge oxygen free radicals and (e) attempts to prevent lung fibrosis. This aspect of paraquat toxicity has been reviewed by others see Meredith and Vale, (1995); Jaeger et al., (1995) Dinis-Oliveira et al., (2008); and section 1.3 on human poisoning in this review.

83.2.23  Absorption from the   Gastrointestinal Tract As discussed earlier, paraquat is poorly absorbed from the gastrointestinal tract and therefore attempts to reduce its entry into the blood stream could be beneficial. Peak blood levels are detected within 60–90 min in rats, dogs and monkeys (Figures 83.1 and 83.2), therefore any interventions must be taken quickly after poisoning if they are to be effective. The bipyridilium herbicides have been shown to bind very strongly to soil and clay minerals (Knight and Tomlinson, 1967). Clark (1971) demonstrated that bentonite and Fuller’s earth where able to reduce mortality in rats given a lethal dose of paraquat when delayed for 2 or 3 hours after paraquat administration. Smith et al. (1974) subsequently showed that repeated doses of a bentonite, castor oil and magnesium sulphate mixture protected rats against the lethal effects of paraquat when given 4 hr after exposure and this regimen also reduced mortality when delayed for as long as 10 hr after exposure. The basis for the protection was shown to be due to a reduction in the concentration of paraquat in the bloodstream and a concomitant reduction in the amount accumulated into the lung

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(Smith et al., 1974). Several other absorbent or binding agents have been examined, activated charcoal was shown to be very effective in rats (Okonek et al., 1982) and in mice in combination with magnesium citrate, the magnesium salt affording some protection on its own (Gaudreault et al., 1985). Kayexalate (sodium polystyrene sulphate), Kalimate (calcium polystyrene sulphate), sodium dextrin sulphate, sodium glucose sulphate and a variety of alkylsulphates and alkylsulphonates have been shown to afford some protection in rats and mice (Nokata et al., 1984; Ukai et al., 1987; Tsuchiya et al., 1995). More recently alginate was found to reduce the absorption of paraquat and hence blood levels in the rabbit (Heylings et al., 2007, see also Section 1.3).

83.2.24  Removal from the Bloodstream Both peritoneal dialysis and haemodialysis have been suggested for removing paraquat from the blood stream and thereby reducing availability to the lungs. Charcoal haemoperfusion was initially demonstrated to remove paraquat from the blood of beagle dogs (Maini and Winchester, 1975). Haemoperfusion appeared to reduce mortality in dogs when given within 12 h of administration of paraquat (Widdop et al., 1975), although more recent studies in the dog have indicated that unless started within 2 hours of exposure it is unlikely to reduce the paraquat content in the lungs (Pond et al., 1993).

83.2.25  Prevention of Accumulation   into the Lung Paraquat is actively transported into alveolar type I and II cells where it accumulates. Studies in vitro using polyamines, diaminoalkanes and a number of other chemicals have identified chemicals that can reduce paraquat accumulation (Lock et al., 1976; Maling et al., 1978; Ross and Krieger, 1981; Smith and Wyatt, 1981; Gordonsmith et al., 1983). However, attempts to reduce paraquat mortality in rats with these agents have failed to demonstrate significant protection (Maling et al., 1978; Dunbar et al., 1988). Another approach has been to use antibodies to paraquat (polyclonal, monoclonal or specific Fab fragments) to try and reduce toxicity to the lung. This approach has been shown to reduce paraquat uptake and cytotoxicity in rat lung slices and isolated alveolar type II cells (Wright et al., 1987a; Chen et al., 1994). However, treatment of paraquatintoxicated mice (Wright et al., 1987b; Cadot et al., 1985) or rats (Nagao et al., 1989) by immunotherapy did not reduce the concentration of paraquat in the lung or affect the mortality.

83.2.26  Free Radical Scavenging Once inside a cell paraquat can redox cycle and produce superoxide anion, singlet oxygen and hydroxyl radicals.

Hayes’ Handbook of Pesticide Toxicology

Many studies have been aimed at attempting to scavenge the radicals formed to reduce or protect the lung injury. In many of these cases significant protection can be demonstrated using isolated cell systems, but in whole animals the protection is limited or equivocal. Superoxide dismutase has been reported to increase survival in rats exposed to paraquat (Autor, 1974; Wasserman and Block, 1978), while other studies have failed to confirm these observations (Frank, 1983; Patterson and Rhodes, 1982). The short plasma half -life of exogenous superoxide dismutase and the fact that it does not enter cells accounts for the lack of protection. A more recent report indicated that a low molecular weight metalloporphyrin superoxide dismutase mimetic afforded some protection against paraquat-induced injury to the lung, but its effect on mortality was not examined and its effect is likely to have been marginal (Day and Crapo, 1996). Desferrioxamine (DF) is an iron chelating agent which has been used to scavenge free iron and thereby reduce hydroxyl radical production. Studies in mice suggested that DF given 24 h before, and regularly after an acute dose of paraquat reduced mortality (Kohen and Chevion, 1985). In this same model these workers showed that iron increased paraquat toxicity. In rats however, DF appeared to afford no protection (Osheroff et al., 1985; Hoffer et al., 1992). Van Asbeck et al. (1989) gave DF by continuous infusion to vitamin E deficient rats and showed it prevented the lung injury and hence reduced mortality. This group also examined the effect of DF and CP51 an hydroxypyridin4-one iron chelator in rats with a normal vitamin E status and found no protection with DF while CP51 increased survival (Van der Wal et al., 1992). Xanthine oxidase inhibitors may also reduce superoxide anion formation and rats fed a diet rich in tungstenate showed a better survival following paraquat exposure than rats fed the diet alone (Kitazawa et al., 1991). Clofibrate induces hepatic peroxisomes in rodents and thereby increases hepatic catalase activity, and it was postulated that a similar effect in the lung might afford protection against paraquat toxicity. Prior administration of clofibrate to rats for 6 days followed by paraquat afforded significant protection, however when clofibrate was administered after paraquat it gave no protection (Frank et al., 1982). Vitamin E is a lipid soluble antioxidant and radical scavenger, some early studies showed that vitamin E deficient animals were more susceptible to paraquat than those with a normal vitamin E status (Bus et al., 1975; Block, 1979). Acute administration of vitamin E to normal mice or rats did not however significantly protect against the toxicity (Bus et al., 1976a; Redetzki et al., 1980) even when instilled into the trachea in a liposome either alone or in combination with reduced glutathione (Suntres and Shek, 1995, 1996). The protective effect of selenium has been reported; animals fed selenium deficient diets being more sensitive

Chapter | 83  Paraquat

to paraquat toxicity (Cagen and Gibson, 1977; Omaye et al., 1978). This is probably related to the seleniumdependant enzyme glutathione peroxidase which plays an important role in protecting cells against oxidative stress. Evidence that glutathione peroxidase plays a key role in protecting animals against paraquat toxicity comes from recent studies in transgenic mice where deletion of this enzyme enhances toxicity while addition affords some protection (Cheng et al., 1998; De Haan et al., 1998). Vitamin C a water soluble antioxidant has provided equivocal data with one study suggesting it might protect while others showed it either had no effect or enhanced paraquat toxicity (Matkovics et al., 1980; McArn et al., 1980; Montgomery et al., 1982; Sullivan and Montgomery, 1984; Minakata et al., 1996). A combination of vitamin C and riboflavin in rats produced a significant improvement in paraquat mortality (Schvartsman et al., 1984), while vitamin C or riboflavin alone was not protective. These authors suggested that perhaps the combination of antioxidant plus an effect of riboflavin on glutathione reductase activity may have contributed to the protection. Niacin has been reported to modestly reduce paraquat mortality in rats, this may be due to an effect of niacin on NAD synthesis which is reduced by paraquat (Brown et al., 1981), however subsequent studies were unable to confirm any protection with niacin (Hooper et al., 1983). Recently sodium salicylate and its pro-drug lysine N-acetylsalicylate have been shown to provide very good protection against the mortality produced by paraquat in rats given either 25 mg/kg ip MLD or 125 mg/kg orally an MLD when give 2 h after paraquat. The authors report complete survival after 30 days compared with no survival in the untreated group after day 5 (Dinis-Oliveira et al., 2007; 2009). Sodium salicylate has anti-inflammatory propertes but most importantly it can form charge transfer complexes with paraquat and thereby prevent oxidative stress. A number of sulphydryl compounds have been examined based on their antioxidant ability, and on an early observation by Bus et al. (1976b) showing that diethylmaleate, which depletes glutathione, enhanced paraquat toxicity. In general precursors of glutathione synthesis which increase intracellular cysteine content have been shown by some workers to provide some increased survival in mice or rats, while others have found these reagents to produce equivocal effects. The protection may be due to alteration of the pharmacokinetics of paraquat or induction of some of the enzymes involved in providing protection against free radical damage. The following have been examined N-acetylcysteine (Shum et al., 1982; Cramp, 1985; Wegener et al., 1988; Hoffer et al., 1993; Hybertson et al., 1995); glutathione (Matkovics et al., 1980; Szabo et al., 1986); cysteine and cystine (Szabo et al., 1986; Kojima et al., 1992); L-2-oxothiazolidine-4-carboxylate (Ali et al., 1996); D-penicillamine (Szabo et al., 1986) and sulphite or thiosulphate (Yamamoto, 1993).

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The affect of the lung-surfactant stimulating drug, ambroxol, has been examined in rats and shown to increase the rate of survival after paraquat (Salmona et al., 1992) while Nemery et al. (1992) found no protective effect.

83.2.27  Prevention of Lung Fibrosis Since delayed deaths with pulmonary fibrosis are a characteristic of paraquat poisoning in experimental animals and humans (see later) a number of agents have been examined to try and ameliorate the fibrotic response. Immunosuppressants such as methylprednisolone, dexamethasone and cyclophosphamide have been examined in experimental animals and in general they were either without effect (Seidenfeld, 1985) or only afforded some protection when give prior to paraquat, but not when given simultaneously (Reddy et al., 1976; Smith and Watson, 1987; Kitazawa et al., 1988). In contrast, Dinis-Oliveira et al. (2006a) reported increased survival of rats given a MLD of paraquat followed 2 h later by dexamethasone. See later for the current position on immunosuppressant drugs in humans. Lung irradiation (Saenghirunvattana et al., 1992) and collagen synthesis inhibitors such as D, L-3,4-dehydroproline (Akahori and Oehme, 1983) were not effective in reducing paraquat lung damage. However, intrathecal administration of an antisense oligonucleotide to heat shock protein 47, a collagen-specific molecular chaperone that assists in the post-translational modification of procollagens during collagen biosynthesis, has been shown to reduce lung fibrosis in rats given paraquat (Hagiwara et al., 2007). One suggestion has been mechanical ventilation with additional inhalation of nitric oxide, based on nitric oxide’s vasodilatory effect on the lungs (Berisha et al., 1994). This approach has not been examined in experimental animals but some clinical experience in combination with other antidotes has been examined (see later). In summary, removal of the ingested material by emesis and purgation of the gastrointestinal tract is currently the most effective method after paraquat exposure in experimental animals. As discussed later, a cocktail of many of these approaches is often used in cases of human poisonings.

83.3  Toxicity to Humans 83.3.1  Experimental Exposure The percutaneous absorption of radiolabelled paraquat has been determined in humans (Wester et al., 1984). Following application of 9 g/cm2 the amount absorbed was 0.29% for the leg, 0.23% for the hand, and 0.29% for the forearm. This gave a calculated in vivo absorption rate of 0.03 g/cm2 for the 24 hr exposure period. Paraquat was thus only minimally absorbed, especially in comparison with other commonly available pesticides (Wester and Maibach, 1985).

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83.3.2  Accidental and Intentional   Poisoning The first case fatalities described involved accidental ingestion of the 20% paraquat concentrate (Bullivant, 1966, Swan, 1967, Oreopoulos et al., 1968, Campbell, 1968). A major source of poisoning was the decanting into unlabelled drinks bottles and other containers (Malone et al., 1971). Throughout the 1970’s the number of reported cases continued to rise, however, there was a noticeable shift in the circumstances. For example, in the Republic of Ireland the number of accidents due to decanting decreased between 1967 and 1977 from 45% to 4% of total cases (Fitzgerald et al., 1978b). Further analysis of the circumstances of poisoning showed that before 1975 there was an approximately equal proportion of accidental and suicidal cases, whereas after that date suicides accounted for over 90% of cases and all fatalities. A similar pattern was described in Northern Ireland (Carson and Carson, 1976) and the United Kingdom (Howard, 1979a, Bramley and Hart, 1983). A review of deaths from pesticide poisoning in the United Kingdom between 1945 and 1989 showed that the number of paraquat-associated deaths rose continuously from 1973 onwards and peaked in 1981 (Casey and Vale, 1994). Since then, the number of deaths has steadily declined to pre-1973 levels. The number of paraquatrelated enquiries to the National Poisons Information Service (London) represented 0.6% of the call load in 1980 but only around 0.05% from the mid 1990’s onwards (Northall and Wilks, 2001). With the increasing use of paraquat throughout the world it became apparent that the problem of intentional paraquat poisoning had shifted away from the British Isles and Europe (Onyon and Volans, 1987). In the 1980’s a high incidence was reported in particular from Asian countries (Naito and Yamashita, 1987, Hettiarachchi and Kodithuwakku, 1989, Goundar, 1984) and the Caribbean (Hutchinson et al., 1991, Perriens et al., 1989). Newer data are sparse; while paraquat self-poisoning continues to be reported from countries such as Korea (Lee et al., 2009c), Sri Lanka (Dawson and Buckley, 2007) and Surinam (Graafsma et al., 2006), worldwide it appears to be only a small fraction of self harm with pesticides which is predominantly related to ingestion of cholinesterase inhibitors (Gunnell et al., 2007). Paraquat poisoning is uncommon in the USA, although it is a large market for paraquat-containing products. A ten year survey of calls to US poison centres showed that paraquat (and diquat)-related enquiries accounted for only around 0.01% of the total (Hall, 1995). Most cases showed either no or minor symptoms, with less than 2 fatalities occurring annually, almost all of them related to suicides. Data on mortality from paraquat poisoning are difficult to compare because of differences in circumstances, treatment and reporting systems. In a collection of data from

Hayes’ Handbook of Pesticide Toxicology

14 publications compiled by the International Programme on Chemical Safety (IPCS, 1984), mortality ranged from 36% to 100%, with an overall mortality of 48% (446 of 925 cases). A difference in mortality between ingestion of the liquid concentrate (20% paraquat ion) and a granular product (2.5% paraquat, 2.5% diquat) has been described by some authors. Park et al. (1975) found that the fatality rate was 15 of 23 (65%) in patients who had ingested liquid concentrate and 3 of 8 (38%) in patients ingesting the granular product. Fitzgerald and Barniville (1978) reported no deaths in 14 patients ingesting the granular product compared to a mortality of 74% in 118 cases of ingestion of the liquid concentrate. In the series published by Howard (1979a) there were 36 deaths from 41 cases (88%) where liquid concentrate was ingested and 5 deaths from 27 cases (19%) involving the granular product. These differences are largely a reflection of the size of dose ingested. While suicidal ingestion of paraquat concentrate accounts for most of the recorded fatalities, the problem of accidental ingestion prompted the principle manufacturer of paraquat to introduce formulation changes to the liquid concentrate in the late 1970’s and early 1980’s (Sabapathy, 1995). A blue colour was added to prevent confusion with drinks, a stenching agent was introduced to alert users, and an emetic was included. In addition, packaging and labelling was improved to prevent decanting of the product, and education and training efforts were directed in particular towards smallholder farmers in developing countries, where the majority of incidents occurred. The effect of these efforts is believed to have made a significant contribution to the decrease of accidental paraquat ingestion in many countries (Sabapathy, 1995, Wesseling et al., 1997). More recently, initial findings from Sri Lanka have suggested that the introduction of a paraquat formulation containing an alginate, magnesium sulphate and an increased level of emetic was followed by a moderate reduction in mortality following paraquat ingestion (Wilks et al., 2008). These results await further confirmation. Although ingestion is the route of entry into the body for the overwhelming majority of poisoning cases, there are a few reports of systemic effects from inhalation and dermal exposure (localised skin, eye and upper respiratory effects will be discussed under ‘Use Experience’). Inhalation exposure is not a prominent feature in paraquat poisoning cases because of the extremely low (not measurable) vapor pressure of paraquat. Respiratory exposure to paraquat during spray applications is very low because the large droplet size will prevent the material from going beyond the nasal cavity. Concerns about oral exposure to spray droplets as a result of drainage into the oral cavity and swallowing appear unwarranted because the typical spray concentration of paraquat for hand-held spray applications is 0.1–0.2%; and would thus require a dose of 1–2 litres of spray solution directly into the nose and into the oral cavity to achieve an lethal dose (Howard, 1980). It is therefore not surprising

Chapter | 83  Paraquat

that there are no reports in the published literature of deaths arising from confirmed inhalation exposure. A review of 30 cases of presumed inhalation exposure found no evidence for systemic poisoning (Vlachos and Kontoes, 1987). Where paraquat was measured it was undetectable or at the limit of detection. Patients were either asymptomatic or had non-specific symptoms such as headache, nausea or feeling unwell. Two patients described nose bleeds. In two patients who presented with cough and fever, pneumonia was established as clinical diagnosis. Reviews of reported cases of possible respiratory exposure (Garnier, 1995, Garnier et al., 2003) concluded that there was only one convincing reported case of possible systemic poisoning following inhalational exposure to paraquat and signs of toxicity were very mild and the patient made a full recovery (Malone et al., 1971). In this case, a 43 year old market gardener sprayed a ‘stronger than usual’ solution (no details of spray concentration available) in a greenhouse and complained of a burning sensation in throat and mouth and weakness. There was biochemical evidence of mild renal failure, but liver function tests and chest x-ray were normal. Paraquat tested positive in urine. Renal function parameters returned to normal within 10 days after exposure. In two fatal cases suggested to be due to respiratory exposure (Wesseling et al., 1997) the possibility of ingestion could not be formally excluded. It has already been mentioned that paraquat absorption across intact human skin is extremely low both in vitro (Walker et al., 1983) and in vivo (Wester et al., 1984). In fifteen cases of single exposures of the skin and eyes during work with paraquat solutions only localised lesions (dermatitis, vesicles, burns, and conjunctivitis) were found (Hoffer and Taitelman, 1989). Paraquat was undetectable in plasma except for three cases where it was at the limit of detection. There were no manifestations of systemic toxicity. A small number of case reports describe systemic paraquat poisoning and fatalities from dermal exposure. In six cases there was deliberate or accidental application of paraquat concentrate to the skin, usually in the unfortunate mistaken belief that it could act against parasitic disease (Ongom et al.,1974, Binns, 1976, Wohlfahrt, 1982 (2 cases), Tungsanga et al., 1983, Garnier et al., 1994). Three cases (Waight, 1979, Okonek et al., 1983, Wesseling et al., 1997) involved widespread accidental contamination of the lower abdomen and legs with the 20% concentrate. In two cases (Jaros et al., 1978, Levin et al., 1979) it was evident that a far too concentrated paraquat dilution (28 g/l; 2.8% and 40 g/l; 4%, respectively) was applied combined with faulty leaking spray equipment and lack of skin decontamination. In a further case (Athanaselis et al., 1983) it is explicitly claimed that a correct dilution of 0.5% paraquat was used (the maximum recommended rate for knapsack). However, subsequent investigation (Hart, 1984) led to the conclusion that, in fact, a more concentrated paraquat solution, probably in excess of 1.5%, was used.

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In one case (Fitzgerald et al., 1978a) the combination of paraquat exposure and pre-existing skin disease caused the death of the person involved, although very few details are given. Another case (Garnier et al., 1994) involved the application of multiple herbicidal mixtures, including paraquat, over several days by a man with a history of psoriasis. This man suffered a febrile lung disease but made a complete recovery. Five cases involve prolonged skin contact with ‘diluted’ paraquat without pre-existing skin lesions being mentioned. The two cases described by Wohlfahrt (1982) give very few data which would be useful in this context. In the third case (Papiris et al., 1995), a farmer was exposed for 5–6 hours to diluted paraquat from a leaking sprayer which caused burning, blisters and erosions in his scrotal area. This patient survived after hospital treatment. In the fourth case (Wesseling et al., 1997), a plantation worker experienced chemical burns on his back, scrotum and inner parts of both thighs after spraying paraquat with a leaking knapsack sprayer for three consecutive days. He subsequently died from interstitial fibrosis of the lung. The fifth case involved an 81 year old man who slept overnight wearing trousers which had been soaked in spray solution and died from renal failure and an acute respiratory distress syndrome after 4 days (Soloukides et al., 2007). No details were given on the concentration of the paraquat solution. Thus, there is no indication that paraquat has caused fatal poisoning through skin contact in normal occupational use. The few cases described in the literature occurred as a result of a combination of factors such as misuse (wrong dilution), pre-existing extensive skin disease, faulty equipment, prolonged extensive skin contact and disregard of safety procedures (no decontamination following significant exposure).

83.3.3  Use Experience Exposure to paraquat under actual field conditions has been assessed in studies with hand held (knapsack), vehicle mounted and aerial applications. Dermal exposure was measured either in patches placed on different body regions or, more recently, using whole body exposure assessments. Inhalation exposure (including oral exposure) was determined using personal air sampling and the air concentration of different particle sizes was measured. Internal dose was assessed using biological monitoring, for which paraquat is an ideal candidate: it is not metabolised, rapidly and completely excreted via the kidneys, stable in urine, and there are sensitive analytical techniques available. The data from these studies are summarised in Table 83.5. There is an enormous variation in dermal exposure evident in the studies found in the literature. This is not surprising given the differences in spray strength, volume applied, application technique, environmental conditions, use of personal protective equipment, and differences in

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Table 83.5  Worker Exposure and Absorption of Paraquat Reference

Country

Application Method

Spray Dilution (%w/v)

Dermal Exposure (mg/hr)

Inhalation Exposure (mg/hr)

Urine level (mg/l)

Swan (1969)

Malaysia

Hand held

0.05

-

-

0.01������ –����� 0.32

Hogarty (1976)

Ireland

Hand held

-

-

0.003

ND

Staiff et al. (1975)

USA

Vehicle mounted

0.1

0.01����� –���� 3.4a

0–0.002

0.02

Hand held

0.2

0.01–0.57a

0.001

0.02

0–0.005

0.05–0.76

0–0.07

0.02–0.03

0.1–2.4b

0–0.047

-

0.05–0.26b#

0–0.06#

0.94–2.71b$

-

0.03

0–0.043

0.03–0.24

0.007

-

Chester & Woollen (1981)

Malaysia

Hand held

0.1–0.2

0.01–12

a

12–170b Wojeck et al. (1983)

USA

Vehicle mounted

0.05–0.1

7.0–42a b

12–169 Chester & Ward (1984)

Chester et al. (1993)

USA

Sri Lanka

Aerial

Hand held

0.3

0.03–0.04

Van Wendel de Joode et al. (1996)

Costa Rica

Hand held

0.1–0.2

0.2–5.7

Singmaster & Liu (1998)

Puerto Rico

Hand held

0.1

-

a

ND Not detected Aerial - flagger# Aerial - pilot$ mg/g paraquat sprayed & extrapolation from indirect measurement using copper as marker. a Exposure to uncovered skin. b Total Dermal exposure.

study design. Nevertheless, some patterns emerge across the variety of study conditions encountered. It is evident that skin exposure represents by far the most significant route of exposure for paraquat. For hand held applications, total dermal exposure was more than an order of magnitude higher than exposure to uncovered body parts (Chester and Woollen, 1981, Van Wendel de Joode et al., 1996). A similar difference was seen for vehicle-mounted spray applications (Staiff et al., 1975, Wojeck et al., 1983). The lowest dermal exposure was seen for pilots applying paraquat (Chester and Ward, 1984), whereas the total dermal exposure of flaggers is comparable to exposure of uncovered body parts in other spray applications. Inhalation exposure was approximately three orders of magnitude lower than skin exposure (Staiff et al., 1975, Chester and Woollen, 1981, Forbess et al., 1982, Wojeck et al., 1983, Chester and Ward, 1984, Van Wendel de Joode et al., 1996, Singmaster and Liu, 1998). Paraquat proved to be below the limit of detection in most samples. Furthermore, the inhalation potential of respirable droplets was found to be negligible since no respirable paraquat could be measured in the breathing zone of exposed workers (Chester and Ward, 1984). Even under difficult spraying conditions (heavy exertion while spraying on hillsides) paraquat was below the limit of detection (Singmaster and

Liu, 1998). In a recent study from Costa Rica, inhalable dust and airborne paraquat levels were significantly lower than the occupational exposure standard (Lee et al., 2009a). Paraquat is an ideal candidate for biological monitoring because it is excreted unchanged in urine, where it is comparatively stable. Most of the worker exposure studies mentioned above included measurement of paraquat in urine. Overall, the paraquat concentration in urine was low with the majority of samples being below the limit of detection. None of the samples contained paraquat at levels which would be indicative of a risk of poisoning (see below). Topical effects from contact with paraquat during spray operations can occur due to a delayed caustic action of paraquat as a result of poor working practice and hygiene (Howard, 1980). Discoloration (white bands), paronychia and partial or complete loss of nails has been described following contact with concentrated (Samman and Johnston, 1969) and prolonged exposure to diluted paraquat solutions (Hearn and Keir, 1971). Upon cessation of exposure, normal nail growth resumes. Irritant dermatitis, burns and blistering can occur from skin exposure to paraquat concentrate or as a result of prolonged skin contact with contaminated clothing or from leaking spray equipment (Swan, 1969, Van Wendel de Joode, et al., 1996).

Chapter | 83  Paraquat

Epistaxis has been described (Swan, 1969, Van Wendel de Joode, et al., 1996), most likely from breathing in spray mist or contact with contaminated fingers. No serious or long-term effects have been described. There are a number of case reports of eye damage resulting from splashes with paraquat concentrate (Cant and Lewis, 1968, Joyce, 1969, Peyresblanque, 1969, Watanabe et al., 1979, Deveckova and Mydlik, 1980). Apart from eye irritation and blepharitis, more serious, delayed ocular damage may occur such as destruction of the bulbar and tarsal conjunctiva and erosion of the corneal epithelium. Anterior uveitis has also been noted. Progressive keratitis and decreased visual acuity may occur, and ocular inflammation with pseudomembrane formation may persist (Vlahos, et al., 1993). However, with appropriate care even severe cases can have a good outcome (McKeag et al., 2002). Attempts have been made to establish the frequency of topical effects from paraquat exposure, particularly for hand held applications in developing countries. Surveys have been carried out interviewing 400 smallholder farmers using paraquat in Malaysia (Whitaker, 1989a), 365 smallholders in Central America (Whitaker, 1989b) and 732 smallholders in Thailand (Whitaker et al., 1993). These surveys showed that, in general, farmers were aware of the potentially fatal consequences of swallowing small quantities of the concentrate. Spray practices and standards of personal hygiene were generally adequate, although the wider use of gloves and eye protection when handling the concentrate needed to be encouraged. In all three surveys, approximately 10% of respondents had experienced health effects attributed to the use of paraquat. These were predominantly skin irritation (mainly on hands and feet), nausea and headaches associated with the smell of the product (due to the added stenching agent) and, to a lesser extent, eye irritation, nail damage and epistaxis. Ramasamy and Nursiah (1988) interviewed 1219 Malaysian estate workers, rice farmers, vegetable growers and smallholders about health effects from pesticide use. They found that exposure to organophosphorous insecticides was associated with giddiness and nausea, whereas the main effects associated with paraquat exposure were eye irritation, nail damage and nasal bleeding. However, their survey did not establish cause effect relationships with exposure to specific products. Only three cases of hospitalisation were described among their study population. Wesseling et al. (2001) reported on acute pesticide related illness among banana plantation workers in Costa Rica in 1996 and reported an overall rate of 2.6 per 100 workers per year for topical injuries and systemic poisonings. In a recent survey, the incidence rate for incidents requiring hospital treatment among 250 Costa Rican smallholder farmers was similar at 3.2 per 100 (Tomenson and Matthews, 2009). However, whereas Wesseling et al. (2001) reported that paraquat was the pesticide most frequently associated with injuries, only 3 of the 16 Costa Rican farmers who were able to identify

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a product responsible for their incident in the more recent survey cited paraquat as the cause of their agrochemical related incident. The State of California has probably the most comprehensive surveillance system of pesticide-related illness in the world. Between 1971 and 1985 a total of 231 cases of ill health attributed to paraquat were notified to the Worker Health and Safety Branch, California Department of Food and Agriculture (Weinbaum et al., 1995). Of these, 38.5% were listed as systemic effects (mainly dizziness, nausea, lightheadedness, headache, chest pain, vomiting and tiredness), 32% were eye effects (burning, itching, redness), 26% were skin effects (rash and irritation, itching) and 3.5% were local respiratory irritant effects (epistaxis, sore throat). There were no cases of pulmonary fibrosis. Analysis of data from 1981 to 1985 showed that the overall incidence of illness was low at 0.6 per 1000 paraquat applications. Detailed medical surveys have been carried out to determine whether the long term exposure to paraquat leads to chronic health effects in workers and spray applicators. Swan (1969) found no abnormalities in chest radiographs of groups of Malaysian rubber plantation workers during paraquat applications over several weeks. Howard (1979b) studied two groups of paraquat formulation workers in the United Kingdom and Malaysia. Mean exposure duration for the UK workers was 5 years, and 2.3 years for the Malaysian workers. A history of skin rashes was found in half of the Malaysian workers, but not in the UK workers where the most common finding was epistaxis and nail damage. Eye irritation was more common in the Malaysian than in the UK workers. There was no evidence of any long-term or permanent skin or eye damage. Comprehensive medical surveys in paraquat-exposed spray operators were carried out in Malaysia (Howard et al., 1981) and Sri Lanka (Senanayake et al., 1993). In both studies there were detailed clinical examinations, lung function measurements (including CO diffusion capacity), haematological and biochemical investigations, and, in the Sri Lankan study, a chest radiograph was taken. In the Malaysian survey, 27 paraquat spraymen (mean spraying time 5.3 years; mean individual annual quantity of paraquat handled 67.2 kg as paraquat ion) were compared with two control groups comprising 24 general plantation workers and 23 latex factory workers, respectively. In the Sri Lankan survey, 85 paraquat spraymen (mean spraying time 12 years) were compared with two groups of 76 factory workers and 79 general workers, respectively. In both studies there were no clinically significant differences in any of the parameters studies, in particular, the results of the lung function tests showed similar results for exposed and control groups. It was concluded that the long term spraying of paraquat was not associated with any measurable adverse health effects. An investigation from 1987/88 in Nicaragua (CastroGutierrez et al., 1997) looked at a population of 134 spray

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workers with at least 2 years spraying experience with paraquat from 15 banana plantations. Of these, 63 had not experienced skin irritation, and 71 who had a history of skin rash or burn (used as a surrogate measure of intensity of exposure). A questionnaire was used to check for symptoms of respiratory illness and Forced Vital Capacity (FVC) and Forced Expiratory Volume in 1 second (FEV1) were measured. The results were compared with a control population of 152 unexposed workers. There was a difference in male: female ratio between the exposed and unexposed groups (100:34 and 88:64, respectively). Paraquat-exposed workers gave a significantly more frequent history of Grade 3 dyspnea, but not Grade 1 or 2 dyspnea. There was no difference in the occurrence of chronic bronchitis, and episodic dyspnea with wheezing was more frequent in the group with topical effects only. However, there were no differences between exposed and control workers with regard to restrictive (FVC 80% of predicted value) or obstructive (FEV1:FVC 70% of predicted value) spirometry parameters. In fact, the lowest incidence of restrictive changes was found in the ‘intensive exposure’ group. Dalvie et al. (1999) carried out a cross-sectional study with 62 male current herbicide sprayers and 70 male controls in South Africa, using a job exposure matrix as an indirect measure to define paraquat exposure. In key findings the study agreed with previous work, concluding that paraquat exposure showed no effect on respiratory symptoms, spirometry and, crucially, gas transfer. However, the authors focused on a small effect on exercise oxygen desaturation measured by pulse oximetry with the mean desaturation in the highest cumulative exposure group being 3.6% compared to 2.1% in the non-exposed group. Unfortunately, more than 25% of the exercise desaturation traces (32/122) could not be evaluated, calling the robustness of this method under the conditions in which it was used into question. The largest and most recent study involved a total of 338 Costa Rican paraquat handlers and non-handlers from banana, coffee and oil palm farms (Schenker et al., 2004). All subjects underwent detailed lung function testing and subjects 40 years of age and younger completed a full cardiopulmonary exercise test. In linear regression models, cumulative paraquat exposure was not an independent predictor of alveolar volume, carbon monoxide diffusing capacity, peak oxygen uptake, forced vital capacity or oxygen pulse peak. The ventilatory equivalent for CO2, was higher with increased cumulative paraquat exposure but was within the normal range. Overall these findings were consistent with no clinically significant increases in interstitial thickening or restrictive lung disease in this population.

83.3.4  Atypical Cases of Various Origins In a case described by Newhouse, et al. (1978), a farmer’s wife had been spraying paraquat in an orchard for many days. This case is unique in that her complaints started with

Hayes’ Handbook of Pesticide Toxicology

scratches on arms and legs which proved non-healing over four weeks. She was then hospitalized for two weeks and discharged without diagnosis. Two and a half weeks later she was readmitted to the hospital because of increased dyspnoea and wheeziness. She was diagnosed as suffering from systemic arteritis and died 12 days after final admission, some 8 weeks after initial exposure. Although the paper links her disease to paraquat exposure, it is doubtful if paraquat was the cause. Firstly, at no time was paraquat measured in blood or urine. Secondly, the time from exposure to her death was more than eight weeks which is highly unusual for paraquat poisoning. Thirdly, she had a clinical diagnosis (systemic arteritis) which did not include any reference to paraquat poisoning. George and Hedworth-Whitty (1980) attributed a case of non-fatal lung disease to the inhalation of nebulised paraquat. A 64 year old woman noticed spray mist drifting into her garden from a spraying operation in an adjacent field. After some 10 minutes she noticed a chest tightening, and over the next week she became gradually more breathless. She was initially treated with a short course of steroids without much effect. Pulmonary function evaluation some two months later showed severe restriction, but there were no abnormalities in the chest radiograph. She was kept on systemic steroids and her lung function had markedly improved some 7 months after the original incident. Hart (1980) commented that the diagnosis of paraquat-induced lung injury was doubtful. The woman had a history of allergic rhinitis and chronic sinusitis. No previous lung function recording was available and no transfer factor was measured at the time of assessment. The chest radiograph was clear and the description of exposure did not provide convincing arguments for a significant inhalation exposure. In the case described by Katopodis et al. (1993), a 31 year old woman was admitted 4 days after ingestion of 2 g paraquat. The urine test for paraquat was still positive , but her plasma concentration was only 10 g/l. Charcoal haemoperfusion was carried out over the next 5 days, paraquat levels became undetectable in plasma on day 6 and in urine on day 8. The patient survived without evidence of pulmonary involvement. The authors attributed the favourable outcome to the haemoperfusion therapy even at such a late stage after ingestion. However, the low paraquat plasma concentration at the time of admission would have suggested a good chance of survival anyway (see below and Table 83.5). Ragoucy-Sengler and Pileire (1996b) reported a case of paraquat poisoning in an HIV positive patient. Indices of severity of the poisoning suggested a survival probability of 30% on admission, and 3% after 72 hours. The clinical course included acute renal failure and severe hypoxia, however, pulmonary fibrosis did not develop. The patient was discharged with normal pulmonary function 18 days after admission. The authors suggested that the immune deficiency on the basis of the patient’s HIV infection may have prevented the development of pulmonary fibrosis.

Chapter | 83  Paraquat

In a case described by Ernouf et al. (1998) a 47 year old man, while under the influence of alcohol, ingested paraquat which had been decanted into an unmarked a red wine bottle. The patient was a chronic alcoholic. He was admitted to hospital within 3 hours and treated with gastric elimination and antioxidant therapy. Evolution of plasma paraquat concentrations pointed towards a prognosis of delayed death from pulmonary fibrosis, however, the patient died on the fourth day after admission from persistent haemodynamic shock and hypoxaemia. The authors speculated that coningestion of ethanol may have enhanced the toxicity of paraquat through increased absorption from the gastrointestinal tract and/or decreased renal clearance. However, it has also been suggested that alcoholism may have a protective effect against paraquat toxicity on the basis of increased synthesis of superoxide dismutase (Ragoucy-Sengler et al., 1991). Methaemoglobinaemia was described in a patient who ingested ‘Gramonol’, a formulation containing 100 g/l paraquat and 140 g/l monolinuron (Ng et al., 1982). The authors speculated that the superoxide anion and hydrogen peroxide generated by paraquat could oxidise haemoglobin to methaemoglobin. However, in response, Proudfoot (1982) pointed out that monolinuron, along with other substituted urea herbicides, is metabolised to aniline derivatives which are well known methaemoglinaemia and haemolysis causing agents. Furthermore, administration of monolinuron alone had produced methaemoglobinaemia in experimental animals. Instead of a new feature of paraquat poisoning, it appeared therefore that Ng et al. had reported the first human case of monolinuron toxicity. Since then, a further case of paraquat-monolinuron poisoning has been described (Casey et al., 1994) in which the severe methaemoglobinaemia (52%) was successfully treated with methylene blue. However, the patient died after 10 days from the consequences of paraquat poisoning. In 1975 the Government of Mexico began an aerial spraying programme, financed by the United States, to destroy marijuana fields with paraquat. In 1978 analyses showed that 21% of 61 marijuana samples confiscated in California, Arizona and Texas contained paraquat residues between 3 and 2000 ppm (Turner et al., 1978). Further work demonstrated that, nationally, 0.63% of over 100,000 kg marijuana seized contained detectable paraquat levels with a median of 52 ppm (Liddle et al., 1980). Over 70% of the contaminated samples were found in the South West region of the United States, originating almost exclusively from Mexico. Combustion testing suggested that around 0.2% of the paraquat residue would pass unchanged into marijuana smoke (Brine et al., 1981). On the basis of a worst case epidemiological risk assessment it was suggested that some marijuana smokers in the South West region might have been at risk of health effects from paraquat inhalation (Landrigan et al., 1983). However, no clinical cases were identified during these studies. A possible association between paraquat exposure and the development of Parkinson’s disease has been the subject

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of much speculation. The reason for this is, as previously discussed, the apparent structural similarity between paraquat and the synthetic pyridine MPTP which produced severe neuropathies in several dozen drug users in southern California (Langston et al., 1983, Lewin, 1984). The first epidemiological work to draw attention to a possible role of pesticides in Parkinson’s disease was published by Barbeau et al. (1986) who showed that the regional incidence of the illness in Canada was non-uniform and correlated with a genetically determined enzyme deficiency. While there was a correlation between disease incidence and pesticide use, this was also found for industrial areas and wood processing regions. Since then over 40 epidemiological studies on the issue of Parkinson’s disease and pesticide exposure have been published, with varying methodologies and conflicting results (for detailed reviews see Brown et al., 2006, Li et al., 2005). Some studies suggested that the use of herbicides was significantly associated with the development of Parkinson’s disease (Ho et al., 1989, Golbe et al., 1990, Semchuk et al., 1991); others have found no such association (Ohlson and Hogstedt, 1981, Tanner et al., 1989, Koller et al., 1990, Zayed et al., 1990, Tanner et al., 1990). Similarly, the evidence from studies in which paraquat has been mentioned is conflicting, even in studies from the same group. For example, Hertzman et al. published two case control studies, in 1990 and 1994. In the first study (Hertzman et al., 1990), 4 out of 57 cases of Parkinson’s disease but none of the 122 control subjects reported ever handling paraquat. The study had limited power due to the small sample size and lack of details on risk factors. In contrast, the later study (Hertzman et al., 1994) included 127 cases and two control groups of similar size. Both quantity and frequency of exposure were assessed for a number of herbicides/pesticides. The relative risk estimates for paraquat were around 1.1–1.2 according to the type of comparison group (hospital/community), and far from statistical significance. A case-control study of occupational risk factors in Parkinson’s disease conducted in Germany and including 390 cases and a total of 755 population controls (Seidler et al., 1996) found relative risks around 2 for professional herbicide or insecticide users, but only in only 1 case was there reported exposure to paraquat. In a study from Taiwan (Liou et al., 1997), self-reported herbicide and pesticide use was higher in Parkinson’s disease cases (31/120 cases) compared to controls (22/240) leading to a significant excess risk above 3 for paraquat exposure which nevertheless did not differ from that for overall herbicide or pesticide exposure. However, subsequent case control studies studies from Finland (Kuopio et al., 1999) and the USA (Firestone et al., 2005) as well as a cross-sectional study from the USA (Engel et al., 2001) reported nonsignificant relative risks, close to unity or slightly above. The Agricultural Health Study is a cohort investigation of over 79,500 agricultural workers from North Carolina and Iowa.

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For the investigation focusing on self-reported Parkinson’s disease the participants were recruited between 1993 and 1997 and followed up to 1999–2003 (Kamel et al., 2007). Overall, 14 out of a total of 83 prevalent and 11 out of the 78 incident cases of Parkinson’s disease reported paraquat use. The relative risks adjusted for all available confounding factors, were 1.8 (95% CI 1.0–3.4) for prevalent Parkinson’s disease, and 1.0 (95% CI 0.5–1.9) for the more meaningful incident Parkinson’s disease. Two recent case control studies from California have focused on possible joint exposure to paraquat and the fungicide maneb as a risk factor in Parkinson’s disease, using an indirect method to assess residential exposure based on Californian pesticide use records and land use surveys. In the study by Costello et al. (2009) there was no significantly elevated risk for exposure to paraquat or maneb alone, either overall or when subjects were split up according to periods of exposure and/ or age groups. However, the authors found a modestly elevated relative risk (1.75, 95% CI 1.13–2.73) in subjects classified as living in a 500 m radius of an area where both pesticides had been used. Ritz et al. (2009) studied a possible interaction between dopamine transporter variants and pesticide exposure, but provided only data on subjects reported to be exposed to both paraquat and maneb. They found an increased risk related to the number of susceptibility alleles, and with exposure above the median value for both pesticides in subjects with increasing numbers of susceptibility alleles. The main limitation of both studies is the unusual and indirect classification of exposure. Firstly, the California pesticide use record database only records applications of a pesticide for Public Land Survey System grid sections (640 acres), so the authors did not know whether a pesticide was used within a 500 m radius circle (194 acres) around a residence in any year. Secondly, a subject would be categorised as residentially exposed to both products even if the applications did not occur in the same year, or even – in the study by Costello et al. (2009) in only one out of 26 study years. Thirdly, the use patterns for the herbicide paraquat and fungicide maneb in California are very different which makes applications on the same crop in the same area at the same time highly unlikely. Fourthly, the methodology is not suited to identify individual exposure since subjects may or may not have been present at times when applications occurred. The latest case-control study to be published used Parkinson’s disease patients identified through the agricultural insurance system (MSA) in France (Elbaz et al., 2009). Detailed exposure assessments were made through a combination of questionnaire and job exposure techniques, including personal visits to farms. These authors found a modest positive association of Parkinson’s disease with overall pesticide use (odds ratio 1.8, 95% CI 1.1–3.1). This appeared to be predominantly associated with use of organochlorine insecticides. Use of bipyridyls was not associated with an increased risk (odds ratio 1.2, 95% CI 0.6–2.0).

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As discussed above, paraquat and MPTP are two very different chemicals in terms of structure-activity relationships, despite their apparent similarity (Koller, 1986, Miller, 2007). Barbeau’s hypothesis that Parkinson disease patients may be more likely to have a specific hydroxylation defect in the P450 enzyme system which might inhibit their ability to metabolise toxins (Barbeau et al., 1985) does not apply to paraquat because it is not metabolised in mammals. Also, none of the health surveys of paraquat-exposed workers (see above) has revealed any neurological deficits, let alone Parkinson’s disease. Furthermore, there is no evidence of a specific effect of paraquat on the nervous system in acute paraquat poisoning, nor have neurological sequelae been noticed in survivors of paraquat poisoning (Vieregge et al., 1988). Zilker et al. (1988) carried out detailed neurological follow-up examinations in 4 survivors of paraquat poisoning (latency period between ingestion and followup 5–10 years) and 3 patients who had had skin contact with paraquat. It was possible to exclude parkinsonism in all patients. One patient exhibited tardive dyskinesia most likely due to long term therapy with neuroleptic drugs. The authors concluded that acute paraquat exposure does not lead to Parkinson’s disease.

83.3.5  Clinical Findings and Dosage Response Information on the clinical course of paraquat poisoning is mainly based on case reports of patients who swallowed paraquat concentrate with suicidal intent. However, the systemic toxic effects are similar regardless of the route of absorption. Paraquat causes nausea which may be prolonged especially following ingestion of emeticised formulations (Meredith and Vale, 1987), as well as vomiting and diarrhoea as a result of its local irritant effect on the gastrointestinal tract. Patients may develop a burning sensation, soreness and pain in the mouth, throat, chest and abdomen (Vale et al., 1987). Ulceration in the mouth and throat, an inability to swallow saliva, dysphagia and aphonia are common. The presence of buccopharyngeal lesions has no prognostic value (Bismuth et al., 1995), in contrast to oesophageal and, in particular, gastric ulcerations which indicate a poor prognosis (Bismuth et al., 1982). Prominent pharyngeal membranes (‘pseudodiphtheria’) have been reported (Stephens et al., 1981) and perforation of the oesophagus may result in mediastinitis, surgical emphysema and pneumothorax (Ackrill et al., 1978). The further clinical course is dependent on the amount of paraquat absorbed into the body (usually following ingestion). Attempts have been made to quantify the toxic dose from estimates based on the information given by patients. Although such estimates are often unreliable, a consensus has emerged which is based on experience with many patients. This has allowed the identification of three degrees

Chapter | 83  Paraquat

of intoxication which are summarised below (for further details see Vale et al., 1987, and Bismuth et al., 1995).

83.3.5.1  Mild or Subacute Poisoning The smallest fatal dose has been quoted as 16.7 mg/kg (Stevens and Sumner, 1991), however, the original reference (FAO/WHO, 1973) makes clear that this value is erroneously low, since the formulation (‘Weedol’) also contained an equal amount of diquat, so that the total bipyridyl ingestion was approximately 35 mg/kg. This is in line with clinical experience which shows that ingestion of less than 20–30 mg paraquat ion/kg has rarely serious consequences. Patients are either asymptomatic or develop nausea and vomiting. Renal and hepatic lesions are minimal or absent. An initial decrease of the diffusing capacity may be apparent in lung function measurements, but full recovery is the norm.

83.3.5.2  Moderate to Severe Acute Poisoning This occurs following ingestion of more than 20–30, but less than 40 to 50 mg/kg. Apart from the localised lesions described above, patients in this group develop renal failure, usually between the second and fifth day after ingestion. Hepatocellular necrosis may occur. Both these lesions are fully reversible. Delayed development of pulmonary fibrosis is responsible for the generally poor prognosis in this group. Clinically and radiologically this appears around seven days after ingestion, but subtle abnormalities are present much earlier, such as a decreased diffusing capacity. The x-ray often shows patchy infiltration which may progress to opacification in one or both lungs. In thin section computerised tomography, the most common pattern on initial scans is ground-glass attenuation, followed by consolidation with bronchiectasis (Lee et al., 1995). In most cases, pulmonary fibrosis leads to development of refractory hypoxaemia, resulting in death over a period of 5 days to several weeks.

83.3.5.3  Fulminant or Hyperacute   Poisoning In cases of massive ingestion (usually well above 40–55 mg/kg paraquat ion) patients survive less than 4 days and die in cardiogenic shock and multiorgan failure. Apart from renal and hepatic failure, alveolitis and noncardiogenic pulmonary oedema are observed. Other organ systems (adrenal glands, pancreas, heart) are affected and mortality in this group has been suggested to approach 100%. Talbot et al. (1988b) reported a series of nine cases of suicidal paraquat poisoning in pregnant women. In the cases where the outcome was known, one foetus died probably unrelated to paraquat, three died in utero or after delivery but associated with respiratory distress in the mothers, two died in utero (one mother survived and subsequently

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had a normal pregnancy with no evidence of teratogenicity from the previous paraquat intoxication). One foetus was aborted. Previously, Fennelly et al. (1968) had reported the case of a woman who was 28 weeks pregnant and died 20 days after paraquat ingestion. Upon autopsy the foetus showed no abnormalities. A 20 week pregnant patient survived the ingestion of a small dose of paraquat and subsequently delivered a normal child (Musson and Porter, 1982). More recently, two cases have been reported in which two third trimester pregnancies had a successful outcome despite significant maternal paraquat ingestion. Jenq et al. (2005) reported on a 27 week pregnant patient who ingested 40 ml of paraquat concentrate. The urine dithionite test was strongly positive and she was treated with early haemoperfusion and immunesuppressive therapy but developed transient renal function impairment. She spontaneously delivered a healthy baby 3 months after the intoxication. At follow-up after 5 years both mother and child were doing well. In the case reported by Chomchai and Tiavilai (2007), a 17 year old girl who was 36 weeks pregnant ingested an estimated 30 ml of paraquat concentrate. She was treated with gastric lavage and adsorbents and an emergency caesarean section was performed 7 hours after the ingestion. Both mother and infant were treated with immunesuppressive therapy. The mother developed transient mild renal functional impairment but was discharged in good condition after two weeks. After an initial uneventful course, the infant became dyspnoic after 6 days and developed a progressive, diffuse interstitial pattern on his chest x-ray. This improved over the next two weeks and the infant was discharged on home oxygen which was discontinued after 10 months. Despite recurrent episodes of wheezing due to respiratory illness his overall condition had much improved at 16 months. There are now sufficient case reports in the literature to demonstrate that the development of pulmonary lesions is not inevitably fatal. Fitzgerald et al. (1979b) examined 13 survivors of acute paraquat poisoning after a minimum of 1 year. In two children, no clinical, functional or radiological abnormalities were seen. Of the 11 adults, five non-smokers did also show no evidence of pulmonary disease. Four smokers were considered normal on clinical and radiological criteria, but had a mild deficit in pulmonary function which could reasonable be attributed to smoking. Two patients had pronounced arterial hypoxaemia, both having had pre-existing pulmonary disease. In one of these two patients new and persistent infiltrates were seen in radiography which could be ascribed to paraquat lung damage. Hudson et al. (1991) described persistent radiological changes in three survivors of paraquat poisoning. In one case the patient died a year after her first intoxication from a second massive dose of paraquat. Upon autopsy pulmonary changes from the first as well as the second intoxica­ tion were present. Lin et al. (1995) studied 16 survivors of moderate to severe paraquat poisoning after 3 months.

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Detailed lung function showed significant improvements over time. This was confirmed by improvements in chest radiographs which showed some residual interstitial fibrosis, especially in the lower lobes. Bismuth and Hall (1995) reported five cases, all of which had developed a restrictive pulmonary lesion, but who survived. Two patients were followed up for 4 and 10 years, respectively. In the first patient there was an obstructive component to his pulmonary insufficiency (from smoking) which persisted over time. However, the restrictive component gradually improved over several years, with eventual return to near baseline state. In the second patient (a 13 year old adolescent at the time of intoxication) pulmonary function tests were completely normal 10 years after the poisoning. He had also been able to actively participate in sports. In addition to these cases, Bismuth et al. (1996) listed 29 cases from the literature of patients surviving restrictive pulmonary dysfunction following acute paraquat poisoning. Most patients progressively improved over time. Yamashita et al. (2000) presented a long-term follow-up of 12 cases of confirmed paraquat poisoning. Their results indicated that survivors of paraquat poisoning may be left with a restrictive type of pulmonary dysfunction but also showed some recovery over time. More recently, Lee et al. (2009b) reported on 8 patients who had been followed over several years following paraquat ingestion using high-resolution computed tomography and lung function measurements. Both radiological and functional changes showed marked improvement over time, in some cases returning to near normal despite initial extensive changes. The measurement of paraquat plasma concentration has proved to be a reliable indicator of the prognosis of the intoxication. Levitt (1979) was the first to demonstrate a relationship between plasma concentration of paraquat, the estimated time after ingestion and the eventual outcome. Based on results from 79 patients with a reasonably well established time of ingestion, Proudfoot et al. (1979) found that those patients whose plasma paraquat concentration did not exceed 2.0, 0.6, 0.3, 0.16, and 0.1 mg/l at 4, 6, 10, 16,and 24 hours after ingestion survived. This semilogarithmic plot has become known as the predictive line, or ‘Proudfoot’s curve’. Because of the rapidly decreasing plasma concentration in the first few hours following ingestion no accurate prognosis could be given prior to 4 hours. The authors emphasised that the line to separate survivors and non-survivors was meant to be an approximate guide, and the main use should be to help clinicians in the decision which patients needed urgent aggressive treatment. Subsequently, several other methods have been described to establish the prognosis from plasma paraquat concentrations. None of those methods have been found to invalidate the original estimate by Proudfoot et al., but they have added other dimensions which may be of help to clinicians. Scherrmann et al. (1987) used data from 30 patients to extrapolate the predictive line beyond 24 hours

Table 83.6  Predictive Plasma Paraquat Concentrations Beyond 24 Hours Separating Surviving and Nonsurviving Patients (from Scherrmann, 1995) Time (hours)

Plasma paraquat concentration (ng/ml)

24

100

48

86

72

74

96

63

120

54

144

48

168

42

192

37

216

32

240

27

264

23.5

288

20

312

18

up to 15 days after intoxication, this was later modified (Scherrmann, 1995) with data from a total of 52 patients (Table 83.6). The same authors evaluated the relationship between early urine concentrations and clinical prognosis. They also attempted to correlate urine results obtained by radioimmunoassay with those given by the simple colorimetric dithionite test. Data from 75 patients showed a wide variation in urine concentrations within 24 hours of ingestion. All 17 patients with concentrations of less than 1 g/ml survived, whereas 51 out of 58 patients with urine paraquat concentrations of more than 1 g/ml died. No colour was observed in the dithionite test at paraquat concentrations below 0.5 g/ml (Scherrmann et al., 1987, Scherrmann, 1995). Using a sample size of 219 patients, Hart et al., (1984) were able to calculate the probability of survival of the patient from the initial paraquat plasma concentration (Figure 83.6). It was noted that the line denoting a 50% probability of survival correlated well with Proudfoot’s curve. Sawada et al. (1988) categorised their patients into three groups: survivors (n  10), non-survivors who died from respiratory failure (n  9), and non-survivors who died from circulatory failure (n  11). They calculated a severity index of paraquat poisoning (SIPP) from time to treatment since ingestion of paraquat multiplied by the serum level at admission (g/ml). A boundary SIPP of 10 separated survival from death by either cause, whereas a SIPP of 50 separated deaths from respiratory failure

Chapter | 83  Paraquat

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5.5 Percentages denote the probability of survival

5.0

4.0

3.0

2.0 10% 20% 30%

1.0 50% 90%

70%

0 0

4

8

12 16 Hours after swallowing

20

24

28

Figure 83.6  Relationship between the concentration of paraquat in the plasma and the survival of the patient. From Hart et al., 1984, reproduced with permission from The Lancet.

and deaths from circulatory failure. Using data from 128 patients, Ikebuchi et al. (1993) separated survivors and fatal cases by multivariate analysis and established a discriminate function D. Their toxicological index of paraquat (TIP) could then be divided into three types: TIP 1 is characterised by D  0.1 (100% survival probability); TIP 2 has the characteristic 0.1  D  0.1 and here urgent treatment may influence the outcome; in TIP 3 the discriminate function D  0.1, and the probability of a fatal outcome is 100%. Jones et al. (1999) combined data from one treatment centre with those from the literature and plotted log (plasma paraquat concentration) against log(h since ingestion) for 375 patients. The predicted probability of survival for any specified time and concentration was exp (logit)/ [1  exp(logit)], where logit  0.58–2.33  log(plasma paraquat)-1.15  log(h since ingestion). In a recent study from Korea (Gil et al., 2008), data of 375 patients from one treatment centre were analysed for the relationship between paraquat plasma concentration and survival. The upper limit of plasma paraquat concentration in survivors was 2.64 g/ml at 3 hour. All patients with plasma paraquat levels above 3.44 died. The minimum paraquat level of the cases with fatal outcome was 0.12 g/ml at 5 hours, 0.02 g/ml at 12 hours and 0.01 g/ml at 24 hours. The authors suggested that plasma paraquat concentrations were a good predictor of survival but not non-survival at low plasma paraquat levels. This was also the conclusion of Senarathna et al. (2009) who prospectively collected data on 451 patients in 10 hospitals in Sri Lanka to test the published predictive methods of Proudfoot et al. (1979),

Hart et al. (1984), Scherrmann et al. (1987), Sawada et al. (1988) and Jones et al.(1999) to determine if any was superior. They found that all methods showed comparable performance within their range of application, but also that all were better at predicting death than survival. The methods described above depend on the availability of paraquat analysis, and this is often not the case, or at least not in a timely fashion. Investigators have therefore attempted to predict the outcome of the intoxication using biological indices rather than plasma paraquat concentrations. Suzuki et al. (1989) measured the respiratory index (RI) from blood gas analysis and used it as an index of lung oxygenation in 51 patients. Progressive deterioration of the RI above 1.5 was found in 43 non-survivors, whereas the RI remained below 1.5 in the 8 survivors. Furthermore, the time taken from ingestion for the RI to exceed 1.5 was found to be a good indicator for predicting the survival period in fatal cases. The major weakness of this method is that it cannot predict the outcome at the point of first contact with the patient, unlike the methods relying on plasma paraquat analysis. Also, conditions which may influence the RI such as pneumothorax, cardiopulmonary rescuscitation, septic shock, pulmonary oedema and pneumonia limit the usefulness of this method. On the other hand, it can be used at any time after the intoxication, and it is independent from an estimate of time of ingestion. Yamaguchi et al. (1990) reviewed the medical records of 160 patients who had ingested paraquat and calculated an equation derived from serum creatinine and potassium concentrations and arterial blood bicarbonate

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level. When plotted against time of ingestion they were able to estimate the probability of survival in three categories (90%, 38% and 3%). Most recently, a different biological index using creatinine measurement from 18 patients has been proposed by Ragoucy-Sengler and Pileire (1996a). They found that the time evolution of blood creatinine in intoxicated patients was linear during the first 24 hours after admission. The rate of increase of creatinine in the patients with fatal outcome was equal to a constant (zero order kinetics). A rate of creatinine increase over 5 hours (dCreat/dt) of  3 mol/l/h was found in the 12 fatal cases whereas this value remained  1.26 for the survivors. As with the method of Suzuki et al. (1989) this biological index is independent from an estimate of time elapsed since ingestion. It has the advantage that a prognosis can be established within a few hours after admission of the patient using a standard biochemical analysis. However, it is currently based on data from a relatively small number of patients and will thus require further confirmation from a larger dataset. Using multiple logistic regression analysis on data from 602 patients, Lee et al. (2002) studied a range of laboratory parameters as possible predictors of survival after acute paraquat poisoning. They found that the probability of survival increased with blood pH and PaCO2, whereas it decreased with age, respiratory rate, hemoglobin, white blood cell count, bIood urea nitrogen, amylase, and urinary paraquat level, as well as the number of failed organs. Huang et al. (2006) applied the Acute Physiology and Chronic Health Evaluation (APACHE) II system (a score calculated from 12 routine physiological parameters during the first 24 hours after admission) to predict the inhospital mortality of 64 patients with acute paraquat poisoning. Non-survivors had a higher APACHE II score than survivors and the APACHE II system yielded better discriminative power than SIPP, plasma paraquat concentration and estimated ingestion dosage of paraquat.

83.3.6  Laboratory Findings If performed early and serially, pulmonary function tests may be of diagnostic value. However, it has been pointed out that any changes seen are not specific for paraquat poisoning, since they may also occur in other clinical conditions such as pneumonia, pulmonary oedema, pulmonary thromboembolism and advanced degrees of the alveolar capillary block syndrome (Cooke et al., 1973). The abnormalities must be interpreted in conjunction with the clinical picture. As mentioned above, pulmonary function tests in patients with moderate to severe paraquat poisoning are likely to be abnormal much earlier than clinical or radiological findings. A decrease in the carbon monoxide diffusing capacity or transfer factor (DLCO or TLCO) can be noted as early as the first day after intoxication (Baguley et al., 1983). Beginning between the fifth and sixth day there may be a restriction of the FEV1 and the FVC. These

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changes are followed by a drop in the arterial oxygen tension and an increase in the gradient of alveolar to arterial tension. Finally, there is the development of a functional shunt by which a decreasing fraction of the blood passing through the lungs is oxygenated (Cooke et al., 1973). A study by Kao et al. (1999) investigated changes in lung ventilation and alveolar permeability in 13 patients with paraquat intoxication, using 99 mTc diethylenetriamine pentaacetate (DTPA) radioaerosol lung scintigraphy. Those patients (69%) who showed abnormal aleveolar permeability subsequently died. Since serum surfactant protein D (SP-D) is thought to reflect the severity of various lung diseases Gil et al. (2007) studied the correlation of disease severity with SP-D concentrations in 12 patients with acute paraqaut intoxication. There was a significant positive correlation between the SP-D level and PaO2, but there was no difference in initial SP-D levels between survivors and non-survivors. In patients who died within 11–14 days, the extent of lipid peroxidation, expressed as malondialdehyde, was higher than in controls or in patients who survived. Massive doses (death in 1–3 days) did not result in increased levels of malondialdehyde (Yasaka et al., 1981, 1986). Serum superoxide dismutase (SOD) levels were significantly decreased in cases of lethal paraquat poisoning (Nemeth et al., 1985). Better clinical courses were detected if SOD levels were normal or slightly elevated. Extremely increased levels were measured several times in the terminal state and were interpreted as the consequence of liver cell necrosis and intravascular haemolysis. In contrast, Yang et al. (2009) examined the association of total antioxidant status (TAS) in serum to clinical outcome in patients with acute PQ intoxication and found that this was not a significant influence on the clinical outcome. Other laboratory findings, including those reflecting renal and hepatic failure, are non-specific. Detailed renal function studies were performed in three cases of paraquat poisoning who developed acute renal failure (Vaziri et al., 1979). The glomerular filtration rate (estimated by using creatinine clearance) improved for two patients who survived two weeks, illustrating the reversible nature of the renal failure. A mild to moderate transient proteinuria but little albuminuria was observed during the first two weeks after intoxication. Other findings consistent with proximal tubular dysfunction included glucosuria, amino aciduria and increased fractional excretion of phosphorus, sodium and uric acid. Using data from 21 patients with paraquat poisoning, Nakamura et al. (2001) showed that serum concentrations of type IV collagen and tissue inhibitor of metalloproteinase-1 rose over the course of 4 days in nonsurvivors but were not altered in survivors. Many case reports have shown a transient rise in liver enzymes such as ALT and AST, reflecting the centrilobular necrosis and cholestasis often seen at autopsy (Vale et al., 1987). Serum protein was decreased in one case (Bullivant,

Chapter | 83  Paraquat

1966), but increased in another with a large increase in the globulin fractions (Matthew et al., 1968). Peak total serum bilirubin concentration correlated significantly with the alveolar-arterial oxygen difference in a series of 21 patients (Lin et al., 1995). Pancreatic enzymes were analyzed in 34 survivors of acute poisoning and amylase and lipase were found to be elevated in 20 % of patients with a peak at day 7 after intoxication (Gil et al., 2009). The elevation was positively correlated with the plasma paraquat level and creatinine was higher in the elevation group. However, there was no evidence of pancreatitis on CT scanning, and significant abdominal pain was not observed. Normochromic anaemia developed rapidly in five cases reported by Lautenschläger et al. (1974). This was accompanied by suppression of erythropoietin in the bone marrow, but little effect on other aspects of haematopoesis. The bone marrow had returned to normal in one patient who survived and was re-examined 6 months after the intoxication. In the above mentioned study by Lin et al., (1995), the alveolar-arterial oxygen difference also showed a negative correlation with the initial platelet count. Paraquat analysis in plasma and urine has already been mentioned as the key to diagnosis and prognosis of paraquat poisoning. A simple spot test can be performed with urine or gastric aspirate and is based on the reduction of paraquat cation to a blue radical in the presence of alkali and sodium dithionite (Berry and Grove, 1971, Widdop, 1976). These methods can detect concentrations of paraquat in urine down to 1–2 g/ml and may be made semiquantitative if a range of standards are prepared in control samples. Quantitative methods based on the dithionite reaction with a spectrophotometric end-point have also been described to determine paraquat in plasma (Knepil, 1977, Jarvie and Stewart, 1979). An improved spot test using extraction with a silica cartridge has allowed lower detection limits between 0.1–0.5 g/ml (Woollen and Mahler, 1987). The lower limit of detection for paraquat using spectrophotometry following solid phase extraction was 45 ng/ml (Smith et al., 1993). Other methods which have been described include a radioimmunoassay with a sensitivity of 6 ng/ml (Levitt, 1979). Gas chromatography and mass spectroscopy has been used (Draffan et al., 1977) giving a sensitivity of 25 ng/ml. A fluoroimmunoassay achieved a sensitivity of 20 ng/ml (Coxon et al., 1988). Gill et al. (1983) described a high performance liquid chromatography method involving ion-pair extraction on disposable cartridges of octadecyl silica. Most of these methods can be applied to the analysis of plasma, urine and tissue samples.

83.3.7  Absorption No adequate data exist on absorption of paraquat in humans. However, Davies (1987) has pointed out that early estimates of an absorption of less than 5% of an ingested

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dose (Conning et al., 1969) may be an underestimate. He suggested that absorption kinetics in man may be more similar to those seen in the dog, where a rapid, but incomplete paraquat absorption occurs, with peak plasma levels occurring at 75–90 minutes, and almost 40% of the dose absorbed in 6 hours, as judged by the amount excreted in urine (Bennett et al., 1976, Davies et al., 1977). Limited clinical data suggest that having a full stomach may effectively decrease the bioavailability of paraquat (Bismuth et al., 1982, 1995). In humans, the precise time at which the plasma paraquat concentration peaks is unknown. However, paraquat may be detected in urine as early as one hour after ingestion (Meredith and Vale, 1987). To judge by the plasma concentration data published by Proudfoot et al. (1979), peak plasma concentrations in humans are certainly attained within 4 hours. This is in line with the toxicokinetic analy­ sis of data from 18 patients by Houze et al. (1990), who estimated peak plasma concentrations to occur between 2 and 4 hours. However, most patients were admitted to hospital comparatively late, and they could measure peak plasma concentrations in only 2 cases, in both they were seen around 3.5 hours after ingestion.

83.3.8  Distribution The distribution of paraquat appears to be similar in humans and dogs (Davies et al., 1977, Van den Bogaerde et al., 1984), suggesting that the three compartment model described by Hawksworth et al. (1981) (see above) in the dog is also applicable to humans. Smith (1987) pointed out that the concentration of paraquat in plasma in human poisoning cases falls rapidly to much lower levels than described in the rat. In their series, Houze et al. (1990) found that the concentration-time curve in 15 adult patients (not haemodialysed) was best described by a bi-exponential curve, with the elimination half lives of the early and late phase being 5 and 84 hours, respectively. These patients could be divided into three groups: 1. Patients admitted early and having a rapidly fatal course from cardiovascular collapse showed only monoexponential decreases with a mean half life of 7 hours. However, because of the early death of the patients, evaluation of the late phase was precluded. 2. The second group included patients who were admitted early and survived long enough for an evaluation of the late phase. They showed a bi-exponential decrease with mean half lives of 7 and 103 hours, respectively. 3. In the third group hospital admission was delayed and only late paraquat plasma concentrations could be measured. Accordingly, a mono-exponential decrease in plasma paraquat concentrations was observed with a mean half life of 101 hours.

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Acute renal failure occurred in all but one of the patients. The terminal half-life, however, was very long even in the patient with normal renal function, suggesting that the prolonged elimination phase depends not only on renal function but also on the gradual release of paraquat by extravascular tissue into the blood circulation. In six of their cases with fatal outcome, Houze et al. (1990) also determined tissue paraquat concentrations. High concentrations were found in the lungs, kidneys, heart and liver and much lower concentrations in lipophilic organs such as brain and adipose tissue. The apparent volume of distribution ranged from 1.2 to 1.5 l/kg, compared to 2.75 l/kg in the study by Davies et al., (1977). The mean value of the distribution half-life in humans is greater than that reported from animal studies (see above). Assuming a first-order distribution rate constant and an early half-life of 5 hours, paraquat distribution would be achieved within approximately 30 to 40 hours (Houze et al., 1990). The active transport of paraquat into lung tissue in different species, including humans, has been described in detail above. Paraquat accumulation in tissue could be considered as a slow process from a pharmacological point of view, but it is rapid in clinical terms (Bismuth et al., 1987). In a study of the kinetics of paraquat through the heartlung block, Baud et al. (1988) showed that concentrations in the radial artery were usually higher than or equal to the corresponding value in the pulmonary artery. Only one patient who was examined approximately 4 hours after ingestion showed a pulmonary artery concentration clearly higher than that in the radial artery, providing evidence of pulmonary uptake of paraquat. The arteriovenous difference disappeared approximately eight hours after ingestion followed by inversion of this ratio. This suggests that lethal concentrations of paraquat in the lung may be reached less than 10 hours after ingestion. Paraquat crosses the placenta and a case reported by Talbot et al. (1988b) suggests that it is concentrated in the foetus. Following suicidal ingestion of paraquat a premature infant (32 weeks) was delivered by Caesarean section. Both mother and infant died shortly thereafter. Paraquat was measured in maternal blood at 5.6 g/ml and in the infant’s blood at 20.6 g/ml.

83.3.9  Metabolism As in experimental animals, paraquat is not metabolised in humans, but is reduced to an unstable free radical which is then re-oxidised to produce a superoxide radical (see above). Paraquat is excreted unchanged in urine.

83.3.10  Excretion As in experimental animals, paraquat elimination is essentially renal via glomerular filtration with an element of

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tubular secretion (Bismuth et al., 1988). With normal renal function, clearance of paraquat is greater than creatinine clearance, which enables excretion of high concentrations and large amounts of paraquat within the first hours after ingestion (Davies et al., 1977, Scherrmann et al., 1983). However, ingestion of large doses of paraquat causes tubular necrosis with a rapid decrease of glomerular filtration and tubular secretion. In four cases described by Houze et al. (1990), renal paraquat clearance was lower than creatinine clearance, even in a patient with apparently normal creatinine clearance. Urinary and plasma elimination half lives correlated well. Paraquat may be detectable in urine for a long period of time. Beebeejaun et al. (1971) found paraquat excreted in urine until 26 days after ingestion. In the case of a 14 month old boy, Houze et al. (1990) could detect paraquat in urine for up to three months after ingestion, suggesting on-going release of paraquat from a deep body compartment. Small amounts of paraquat have been recovered in bile samples at post-mortem examination, suggesting that a minor enterohepatic cycle may exist in humans (Van Dijck et al., 1975). As in experimental animals, the amount of paraquat excreted in faeces corresponds to 60–70% of the ingested dose in humans. This excretion may be prolonged (Van Dijck et al., 1975).

83.3.11  Pathology Pathological findings upon autopsy in humans fatalities from paraquat poisoning are similar to those seen in experimental animals, in particular the rat (for a detailed review see Smith and Heath, 1976). The lung is the organ showing the most severe changes in paraquat poisoning. Pulmonary pathology has been divided into two phases which correspond with the early and late stages of the clinical signs and symptoms (Smith and Heath, 1975).

83.3.11.1  The Destructive Phase This occurs during the first few days after paraquat poisoning and is rarely seen in human autopsy cases, but has been described in a case where an early biopsy was performed (Toner et al., 1970). It is characterised by swelling of the alveolar epithelium which sloughs off and is thought to be related to early development of pulmonary oedema with congestion and fibrin exudate (Smith and Heath, 1974a). Death due to this pulmonary pathology is rare.

83.3.11.2  The Proliferative Phase This phase is usually seen in patients who survive for longer than 1 week. Pulmonary congestion with interstitial and alveolar oedema continues, sometimes associated with haemorrhage. There is lymphocytic and other inflammatory

Chapter | 83  Paraquat

cell infiltration and occasional proliferation of cells lining the alveolar wall (Bullivant, 1966). The most specific feature is the presence of large quantities of fibroblastic tissue which is perivascular and peribronchial early on, but later more diffuse (Smith and Heath, 1974b). The pulmonary fibrosis is sometimes associated with an early honeycomb appearance of the lung parenchyma, however, in contrast to a true honeycomb lung the cystic air spaces are dilated respiratory bronchioles and their walls consist of fibrosed, collapsed alveoli. Renal pathology is common, but rarely responsible for the death of the patient. Macroscopically, the kidneys are swollen and soft. There is degeneration or necrosis of proximal tubular cells (Bullivant, 1966, Campbell 1968) with nuclear loss and cast formation (Parkinson, 1980). Depending on the time after poisoning, there may be signs of regeneration. While early studies made little mention of liver damage, Mullick et al. (1981) found evidence of cholestasis, usually localised to the centrilobular region in the majority of their 13 autopsy cases. There was cholangiocellular injury involving the small and medium-sized bile ducts in portal areas. The authors hypothesised that paraquat injury to the liver is biphasic with an initial hepatocellular injury followed after 2 days by a cholangiocellular phase. Toxic myocarditis is frequently seen in cases with ingestion of larger amounts of paraquat. Parkinson (1980) described a patchy but widespread polymorphonuclear leucocyte infiltration in the presence of normal myocardial fibres. In some cases adrenal cortical necrosis has been described (Nagi, 1969, Reif and Lewinsohn 1983) in patients who died early after ingestion of paraquat. This lesion was diffuse and involved mainly the zona fasciculata and zona reticularis. Fitzgerald et al. (1977) found adrenal cortical necrosis upon autopsy in 12 of 23 patients. The severity of the lesion appeared dose-related with patients showing complete cortical necrosis after ingestion of higher doses. Brain pathology has been studied in a series of 8 patients (Grant et al., 1980). Changes included generalised oedema, haemorrhages (these two findings being the most consistent changes), glial reactions and meningeal inflammation. The authors suggested that paraquat may damage the cerebral blood vessels. These changes were also seen in a case reported by Hughes (1988) who suggested that, apart from a direct toxicity of paraquat on cerebral blood vessels, the neuronal depletion, myelin breakdown and astrocytic fibrous gliosis seen were a secondary effect due to prolonged anoxia.

83.3.12  Treatment of Poisoning The therapy of paraquat intoxication has focused on three main areas: prevention of absorption from the gastrointestinal tract, enhancement of elimination of paraquat from

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the body, and therapy directed against the mechanisms of toxicity. In addition, there have been attempts to use lung transplantation as a means to overcome the consequences of paraquat lung toxicity.

83.3.12.1  Prevention of Absorption Following the first reports of paraquat poisoning it was suggested that the immediate therapy of paraquat poisoning should be directed towards prevention of absorption from the gastrointestinal tract (Malone et al., 1971). There is little information available on the use of gastric lavage in paraquat poisoned patients. Bismuth et al. (1982) were not able to establish a beneficial effect from gastric lavage in their series of 28 patients. Bramley and Hart (1983) did not find an improved prognosis resulting from the use of gastric lavage in a series of 262 patients. McDonagh and Martin (1970) proposed urgent gastric lavage with a 1% bentonite solution to inactivate paraquat. Following the studies by Clark (1971) who found that bentonite (sodium montmorillonite) and Fuller’s Earth (calcium montmorillonite) had a high adsorption capacity for paraquat, Douglas et al. (1973) reported three cases of survival after paraquat poisoning, two of which had been treated with 7% bentonite as adsorbent. Smith et al. (1974) suggested a treatment regime of repeated administration of cathartics together with large volumes of Fuller’s Earth or bentonite which had been shown to effectively protect rats against an otherwise lethal dose of paraquat. Vale et al. (1977) used this approach together with charcoal haemoperfusion in 10 patients with paraquat poisoning. Only one patient who had the initially lowest plasma paraquat concentration survived, prompting the authors to conclude that the treatment was likely to be of benefit only in less severely poisoned patients. This was also the conclusion of Fitzgerald et al. (1979a) who analysed 62 cases of paraquat poisoning with respect to treatment with Fuller’s Earth and survival. They found that the majority of patients who survived had not taken what was regarded as a lethal dose. Also, death occurred in all patients who had ingested more than 30 ml of the concentrate, irrespective of therapy. In the group of patients who ingested between 5 and 30 ml and who received therapy within 6 hours after ingestion 4 out of 7 survivors and 2 out of 5 non-survivors had received Fuller’s Earth. The authors suggested that Fuller’s Earth may have been of benefit in a few cases who had taken slightly in excess of the lethal dosage, but was unlikely to affect the outcome in the majority of patients with paraquat poisoning. While Fuller’s Earth is still widely used in the firstline treatment of paraquat poisoning, the original claim by Clark (1971) that activated charcoal did not bind paraquat has been disputed. On the basis of in vitro binding studies and in vivo experiments, Okonek et al. (1982) suggested that the use of activated charcoal instead of Fuller’s Earth was equally effective. This has prompted a revision

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of the advice given to medical practitioners in the United Kingdom (Department of Health, 1996) since activated charcoal is more likely to be immediately available in most hospitals and treatment centres. Nakamura et al. (2000) found that particle size of activated charcoal did not influence the amount of paraquat adsorbed in vitro. However, the authors also showed that the smaller the particle size of activated charcoal, the faster PQ was removed due to the larger contact surface area. More recently, Berry et al., (2007) tested the in vitro binding capacity of a range of brands of activated charcoals and found that all brands performed similar to or better than Fuller’s Earth. Other adsorbents such as the cation exchange resin kayexalate have been used (Yamashita et al., 1987) but it is doubtful whether these have any benefit over the use of Fuller’s Earth and activated charcoal. From 1979 onwards a potent emetic, the phosphodiesterase inhibitor PP796, was gradually introduced in all paraquat formulations made by the major manufacturer (Denduyts-Whitehead et al., 1985). It has been shown that following ingestion of emeticised formulations vomiting occurs earlier, is more profuse and prolonged than following ingestion of non-emeticised product (Meredith and Vale, 1987). However, a comparison of data from patients who had ingested paraquat concentrate with or without added emetic failed to show an overall benefit of the emetic on survival rate (Bismuth et al., 1982, Bramley and Hart, 1983, Onyon and Volans, 1987). Nevertheless, the emetic has been retained with the rationale that in particular accidental paraquat ingestions usually involve small quantities of the product, where early gastric emptying could have an effect on the outcome. It can be concluded that there is no clear evidence that gastric emptying and the use of adsorbents have improved the survival of patients with paraquat poisoning. The main reasons for this are the high dose of paraquat ingested by the majority of patients with deliberate ingestion, and the frequent delay in hospital admission. Most authors concede that, on theoretical grounds, therapy designed to prevent absorption of paraquat should be able to help those patients who have a realistic chance of survival. However, clear evidence for this from clinical studies has so far not been obtained,

83.3.12.2  Elimination of Paraquat   from the Body Since the kidney is the primary excretory organ for absorbed paraquat, enhancement of urinary elimination was one of the first therapeutic options considered. Kerr et al., (1968) published the first case report where forced diuresis had been used to treat paraquat poisoning. The exact fluid volume was not given, but their patient’s urine excretion was more than 11 litres over 24 hours. The total urine excretion of paraquat was 46 mg and the patient survived.

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Another patient was treated with a total of 27 litres of fluid over 48 hours (Fennelly et al., 1971). During the course of the forced diuresis he developed seizures, a metabolic alkalosis and electrolyte disturbances, but the therapy was successfully completed. He developed transient mild hepatic and renal failure, but the only sign of pulmonary involvement was a slight temporary reduction in transfer factor. The authors suggested that this was a case of severe poisoning but were unable to attribute his survival to the forced diuresis therapy because the patient had also received immunosuppressive therapy with azathioprin and prednisolone. Bismuth et al. (1982) suggested that forced diuresis per se does not enhance the urinary elimination of paraquat. Nevertheless, they believed the therapy might be of value in the prevention of paraquat-induced renal damage because of a reduction in the tubular concentration of paraquat. However, of the 18 patients with developing renal failure who were treated with frusemide, only one survived despite the fact that diuresis was maintained in nine patients. Removal of paraquat by means of peritoneal dialysis, haemodialysis and haemoperfusion has been advocated to reduce paraquat plasma concentrations and enhance elimination. Of these, dialysis procedures were found to be ineffective (Vale et al., 1977, Bismuth et al., 1982) and the value of charcoal haemoperfusion remains controversial. Experimental haemoperfusion in dogs was able to improve survival (Widdop et al., 1977), but early results in paraquat poisoned patients were disappointing (Vale et al., 1977). In 1979, Okonek and co-workers published a report on the successful treatment of two patients with what they described as ‘continuous haemoperfusion’. Plasma paraquat analysis prior to haemoperfusion indicated a very poor prognosis, but under an aggressive haemoperfusion therapy over several weeks both survived. Subsequently, a further 6 patients were treated with this regime and had a positive outcome (Okonek et al., 1982/83). However, these apparent successes proved to be rare. Hampson and Pond (1988) carried out a meta-analysis of data from 35 cases published in the literature and 7 cases from their own hospital which had sufficient comparative data, as well as details of the haemoperfusion procedure. They showed that none of the patients whose initial plasma paraquat concentration was higher than 3 mg/l survived, regardless of time after ingestion and treatment. Overall, the outcome was in line with predictions and did not appear to be affected by haemoperfusion, single or repeated. The authors concluded that haemoperfusion should only be considered for patients whose initial plasma concentration was below 3 g/l, those, in whom the probability of survival was between 20 and 70%, and those who present within a few hours of ingestion. Subsequently, Böhler et al. (1992) reported a case where the use of continuous arterio-venous haemoperfusion was effective in lowering the plasma paraquat concentration below the limit of detection. However, the patient died on the second day after ingestion from gastrointestinal complications. Suzuki

Chapter | 83  Paraquat

et al. (1993) compared the effect of ‘aggressive’ ( 10 hours in the first 24 after ingestion) vs. ‘conventional’ ( 10 hours) haemoperfusion on the outcome of the intoxication in 40 patients. Aggressive haemoperfusion did not improve the overall outcome, but significantly increased survival time. Lee and Lee (1995) found that 8 out of 18 patients treated with haemoperfusion survived, whereas none of 20 who did not receive haemoperfusion died. No plasma paraquat concentrations were measured, but the authors stated that the estimated volume ingested was not significantly different between the two groups. Koo et al. (2002) evaluated the effect of prophylactic continuous venovenous haemofiltration (CVVH) in 80 patients with paraquat poisoning who were either treated with haemoperfusion alone (n  44) or a haemoperfusion-CVVH combination (n  36). Prophylactic CVVH after haemoperfusion prevented early death caused by circulatory collapse and prolonged survival time. However, it could not prevent late death caused by respiratory failure and did not provide a survival benefit in acute paraquat poisoning. In the largest case series published so far, Hong et al. (2003) performed haemoperfusion on 105 patients in order to assess the extracorporeal elimination. Paraquat concentration in plasma was measured before and after four hours of haemoperfusion. Since the reduction rate was significantly higher in the survivor group than in the non-survivor group, they concluded that adequate haemoperfusion appeared to be helpful in the treatment of acute PQ poisoning. In conclusion, no consistent benefit has been demonstrated from therapies aimed at enhancing elimination of paraquat from the body. The best chances appear to lie in the maintenance of renal function through adequate diuresis. As for extracorporeal elimination, haemoperfusion appears the only technique which may be of benefit in some patients, and the early and aggressive use of this technique may have contributed to survival in some cases. In their proposed treatment scheme for acute paraquat poisoning, Dinis-Oliveira et al. (2008) suggested to use up to 7 haemoperfusion sessions of 6–8 hours duration, preferably started within 4 hours of ingestion and maintained until plasma paraquat levels would be  0.2 mg/L.

83.3.12.3  Pathophysiological Treatment A wide range of therapeutic substances have been studied experimentally to try and prevent the specific lung toxicity of paraquat from occurring. Some have been used in humans, but most of the published work is based on single or a small number of cases. Usually, more than one therapy was employed, and information on the severity of poisoning and the initial probability of survival is often limited. For these reasons a critical evaluation of the benefit of any one therapy is difficult and, in many cases, impossible. Since oxygen is required to set off the biochemical cascade of paraquat toxicity the use of supplementary oxygen

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should be avoided as long as possible. Bismuth et al. (1982) used a hypoxic breathing mixture and hypothermia in six patients. The arterial oxygen tension was maintained below 6.6 kPa. Only one patient survived who had clinical evidence of only mild poisoning. In the other patients, the FiO2 had to be increased on a daily basis, all of them requiring 0.5 (50%) prior to their death. Since redox cycling and the generation of free radicals are considered to be the principle steps in the development of alveolar epithelial cell damage, a number of agents which, at least theoretically, interfere with this process have been tried therapeutically. One of the first steps in the biochemical cascade of injury is the generation of the superoxide anion which is detoxified by the enzyme superoxide dismutase. This has been given either intravenously (Davies and Conolly, 1975), intramuscularly (Harley et al., 1977), intrapulmonary during fibreoptic bronchoscopy (Bateman, 1987) or as a nebulized aerosol (Davies and Conolly, 1975, Hong et al., 1996). In some cases there was co-administration with the antioxidants vitamin C (Hong et al., 1996) or vitamin E (Harley et al., 1977) which has also been given on its own (Shahar et al., 1980). The doses given appeared to have been determined empirically, and no conclusive evidence of a beneficial effect has so far been shown. Hong et al. (2002) gave increasing doses of vitamin C to 10 paraquat-poisoned patients over 5 days and found an increased total antioxidant status. All patients survived, but the influence of the vitamin C on outcome could not be assessed. N-acetyl cysteine (NAC) is a glutathione precursor which readily crosses the cell membrane, and glutathione depletion is one of the features of paraquat-induced cellular damage. Lheureux et al. (1995) treated a patient with high doses of NAC (300 mg/kg/day) over three weeks. However, the patient who survived also received early haemodialysis and desferrioxamine. The latter, an iron chelating agent, has been proposed because iron has a catalytic effect in the production of hydroxyl radicals. However, no other data exist on its clinical use. Since then, further case reports have included treatment with NAC (Drault et al., 1999, Lopez Lago et al., 2002, DinisOliveira et al., 2006b). These patients also survived, but all received additional treatment with haemoperfusion or haemodialysis, and in the latter case also desferrioxamine, cyclophosphamide and methyl prednisolone. Therefore, no definite conclusion can be drawn on the effectiveness of pathophysiological treatment in the clinical setting, alone or in combination.

83.3.12.4  Prevention of Lung Fibrosis The development of the paraquat lung lesion is characterised by early infiltration of inflammatory cells, followed by fibroblast proliferation. Attempts have therefore been made to halt this process by giving immunesuppressive therapy. A few case reports involved the use of azathioprine, in one

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case with successful outcome (Laithwaite, 1975), in two other cases the patients died (Malcolmson and Beesley, 1975). In one patient who survived, bleomycin was used over three days (Mahieu et al., 1977). Most experience exists with a combination treatment of cyclophosphamide and corticosteroids which was first advocated by Malone et al. (1971). Addo et al., (1984) claimed a 75% survival rate in 20 patients treated with cyclophosphamide (5 mg/kg/day to a maximum total of 4 g) and dexamethasone (8 mg eight-hourly over two weeks). Two years later they published a case series using the same regime with 72 patients, 52 (72%) of which survived (Addo and Poon-King, 1986). However, the plasma paraquat data of 25 patients showed that 7 survivors had no measurable paraquat levels, and of the other 18 only the six patients with the lowest plasma concentration survived. Following a preliminary report (Lin et al., 1996) on the use of pulse therapy with cyclophosphamide (1 g/day over 2 days) and methylprednisolone (1 g/day over 3 days), Lin et al. (1999) reported results of a prospective study in 142 patients. Seventy-one patients who died from fulminant poisoning within one week were retrospectively excluded. In the group of patients classified as moderately to severely poisoned (i.e. those alive after 7 days), only 4/22 patients treated with cyclophosphamide died, compared to 16/28 in the control group. Plasma paraquat concentrations were not available, but the authors stated that there was no difference in severity of poisoning between the two groups based on the urine dithionite test. However, the statistical methodology of the study, in particular the failure to do an intention-to-treat analysis has been criticised (Buckley, 2001). Others have disputed the beneficial effects of the cyclophosphamide-dexamethasone regime (Nogue et al., 1989), and in a prospective study Perriens et al. (1992) did not find any difference in mortality between 14 patients who had received standard treatment and the 33 patients who had received high-dose cyclophosphamide and dexamethasone. In a systematic review on the subject of immunesuppressive therapy in paraquat poisoning, Eddleston et al. (2003) evaluated 10 clinical studies. Mortality in controls and patients varied markedly between studies. Three of the seven non-randomised studies measured plasma paraquat; analysis using Proudfoot’s or Hart’s nomograms did not suggest that immunosuppression increased survival in these studies. The authors concluded that the results could so far be regarded as hypothesisforming rather than conclusive, and that a large, properly randomized and controlled trial was needed. Since then, Chomchai et al. (2004) reported a small case series from one hospital of patients classified as ‘moderately to severely’ poisoned who had been treated with the cyclophosphamide/dexamethasone regime in addition to receiving vitamin B and C. Five of the 6 patients survived, compared to a historical control group of 9 patients who died. However, 7 of these were described as having ‘fulminant’ poisoning. A further study from Lin’s group

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in Taiwan presented results from a prospective randomized trial in patients with a predicted mortality risk of 50–90% using a modification of the immunesuppressive regime employed in earlier studies (Lin et al., 2006). All patients were treated with gastrointestinal decontamination and 2 courses of haemoperfusion over 24 hours. The treatment group patients received an initial dose of cyclophosphamide (15 mg/kg/day over 2 days) and methyl prednisolone (1 g/day over 3 days), followed by dexamethasone (5 mg i.v. every 6 hours) until PaO2 exceeded 80 mmHg. If PaO2 dropped below 60 mmHg, patients would receive an additional course of methyl prednisolone (1 g/day over 3 days). In addition, they would receive a single infusion of cyclophosphamide (15 mg/kg) if at least 2 weeks had passed since the first course and the white cell count was 3,000. The control group received 5 mg dexamethasone every 6 hours until PaO2 was 80 mmHg. Survival was 11/16 patients in the treatment group, and 1/7 patients in the control group. Although these results are again encouraging, they were obtained in a small number of patients; hence, the call for a large, properly randomized and controlled trial is still valid (Eddleston et al., 2003). Because of the radiosensitivity of fibroblasts in vitro, Webb et al., (1984) treated a patient who had developed diffuse alveolar damage following paraquat ingestion initially with cyclophosphamide and, after further deterioration with fractionated radiotherapy over 11 days. The patient survived. It was noted that the severity of poisoning in this patient was mild (Proudfoot et al., 1984) and the majority of patients in subsequent reports died (Bloodworth et al., 1986, Williams and Webb, 1987). This may have been due to differences in the severity of intoxication, as well as the therapy employed. Following the successful treatment of a patient with poor prognosis (Talbot et al., 1988a), Talbot and Barnes (1988) treated a further 8 patients with radiotherapy. Only 2 survived and the authors suggested that a definite benefit of radiotherapy could not be demonstrated in their study.

83.3.12.5  Other Treatments Beta-blocking agents such as propranolol have been shown to block the uptake of paraquat into the lung (Maling et al., 1978). However, their limited therapeutic use has not been successful (Davies and Conolly, 1975, Fairshter et al., 1976, 1979). There have been two case reports on the use of nitrogen oxide inhalation (NO) in paraquat poisoning. On the basis that NO is a potent endogenous vasodilator and that NO inhalation exerts a beneficial effect on pulmonary gas exchange, Köppel et al. (1994) treated a 52 year old patient with severe paraquat poisoning (plasma concentration 4 days after ingestion 1 mg/l). She received 25 ppm in the inhalation mixture, her respiratory parameters improved immediately and she was stabilised for three days.

Chapter | 83  Paraquat

However, the patient died with massive pleural effusions and ventilatory failure on day 11 after ingestion. In the second case, Eisenman et al. (1998) treated a 52 year old male whose plasma paraquat concentration predicted only a 30% chance of survival, with NO because of developing respiratory distress. In addition, the patient had received Fuller’s Earth, forced diuresis, haemofiltration, N-acetyl cysteine, methyl prednisolone, cyclophosphamide, vitamin E, and colchicine. Because of the multiple therapy it was impossible to be sure which of the therapeutic measures had contributed to this patient’s survival. Nevertheless, it was felt that the use of NO deserved further evaluation (Hall, 1998). There are five reports in the literature where lung transplantation has been performed after paraquat poisoning. Matthew et al. (1968) described a single lung transplantation 6 days after accidental paraquat ingestion in a 15 year old boy whose plasma paraquat levels at the time of the operation were still at toxic levels (0.4 g/ml). The patient died 13 days after the operation in respiratory failure and the autopsy showed changes typical for paraquat poisoning, although no paraquat was measurable in the transplanted lung. A contribution of rejection to the disease process could not be excluded. The same group subsequently reported a further unsuccessful lung transplantation in an 18 year old farm worker (Cooke et al., 1973). A further single lung transplantation with fatal outcome in a 25 year old man was reported in 1984 by Kamholz et al. This patient died after 45 days from the consequences of a bronchopleural fistula. Sequential bilateral lung transplantation was described by the Toronto Lung Transplant Group (Saunders et al., 1985). This 31 year old patient had received a lung transplant at a time when his plasma paraquat levels were below what was considered to be a toxic level (0.03 g/ml). In the postoperative period there was a seven-fold increase in paraquat plasma levels, possibly due to the release of paraquat from muscle stores. The transplanted lung failed and following several days of intensive therapy including haemoperfusion a second transplant was undertaken. The transplant worked well over two months, however, the patient developed a progressive toxic myopathy and ultimately died from a cerebrovascular accident. Nevertheless, this case showed the feasibility of lung transplantation in paraquat poisoning. The most recent case (Licker et al., 1998) is the only one with a successful outcome. A 17 year old farmer developed respiratory failure of unknown origin. Repeated plasma paraquat measurements were negative. Following mechanical ventilation for five weeks a single lung transplantation was carried out. Recovery was complicated by myopathy, and paraquat was confirmed in the excised lung and a muscle biopsy. The patient subsequently admitted to having taken paraquat. The patient was discharged after 88 days and was able to lead an independent life at the last follow-up 13 months after transplantation.

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These cases demonstrate that, over the years, lung transplantation has become feasible in cases of paraquat poisoning. While the early attempts were hampered by problems with immune suppression as well as a lack of understanding of the pathophysiological events following paraquat poisoning, these problems appear now to have been satisfactorily resolved. However, the authors of the latest paper make the point that the use of such a scarce and expensive resource is questionable in cases of deliberate self-harm.

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Chapter | 83  Paraquat

Wilks, M. F., Fernando, R., Ariyananda, P. L., Eddleston, M., Berry, D. J., Tomenson, J. A., Buckley, N. A., Jayamanne, S., Gunnell, D., and Dawson, A. (2008). Improvement in survival following paraquat ingestion after introduction of a new formulation in Sri Lanka. PloS. Med. 5, e49. Williams, M. V., and Webb, D. B. (1987). Paraquat lung: is there a role for radiotherapy. Human Toxicol. 6, 75–81. Winterbourn, C. C. and Sutton, H. C. (1984). Hydroxyl radical production from hydrogen peroxide and enzymatically generated paraquat radicals catalytic requirements and oxygen dependence. Arch. Biochem. Biophys. 235, 116–126. Witschi, H., Kacew, S., Hirai, K. I., and Cote, M. G. (1977). In vivo oxidation of reduced nicotinamide adenine dinucleotide phosphate by paraquat and diquat in rat lung. Chem. Biol. Interact. 19, 143–160. Wohlfahrt, D. J. (1982). Fatal paraquat poisonings after skin absorption. Med. J. Aust. 1, 512–513. Wojeck, G. A., Price, J. F., Nigg, H. N, and Stamper, J. H. (1983). Worker exposure to paraquat and diquat. Arch. Environ. Contam. Toxicol. 12, 65–70. Woollen, B. H., and Mahler, J. D. (1987). An improved spot-test for the detection of paraquat and diquat in biological samples. Clin. Chim. Acta. 167, 225–229. Wright, C. E., Tallan, H. H., Lin, Y. Y., and Gaull, G. E. (1986). Taurine: biological update. Ann. Rev. Biochem. 55, 427–453. Wright, A. F., Green, T. P., Robson, R. T., Niewola, Z., Wyatt, I., and Smith, L. L. (1987a). Specific polyclonal and monoclonal antibody prevents paraquat accumulation into rat lung slices. Biochem. Pharmacol. 36, 1325–1331. Wright, A. F., Green, T. P., Daley-Yates, P., and Smith, L. L. (1987b). Monoclonal-antibody does not protect mice from paraquat toxicity. Vet. Human. Toxicol. 29(Suppl 2), 102. Wright, S. H., and Wunz, T. M. (1995). Paraquat 2/H exchange in isolated renal brush-border membrane vesicles. Biochem. Biophys. Acta. 1240, 18–24. Wyatt, I., Doss, A. W., Zavala, D. C., and Smith, L. L. (1981). Intra­ bronchial instillation of paraquat in rats: lung morphology and retention study. Br. J. Ind. Med. 38, 42–48. Wyatt, I., Soames, A. R., Clay, M. F., and Smith, L. L. (1988). The accumulation and localisation of putrescine, spermidine, spermine and paraquat in the rat lung. Biochem. Pharmacol. 37, 1909–1918. Yamada, K., and Fukushima, T. (1993). Mechanism of cytotoxicity of paraquat. II. Organ specificity of paraquat-stimulated lipid peroxidation in the inner membrane of mitochondria. Exp. Toxic. Pathol. 45, 375–380. Yamagami, K., Matsubara, M., Kitazawa, Y., Takeyama, N., Tanaka, T., and Kawamoto, K. (1994). Flow cytometric analysis of the direct toxic effects of paraquat on cultured MDCK cells. J Appl. Toxicol. 14, 155–159.

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Yamaguchi, H., Sato, S., Watanabe, S., and Naito, H. (1990). Pre-embarkment prognostication for acute paraquat poisoning. Human Exp. Toxicol. 9, 381–384. Yamamoto, H. (1993). Protection against paraquat induced toxicity with sulfite or thiosulfate in mice. Toxicology 79, 37–43. Yamamoto, T., Anno, M., and Sato, T. (1987). Effects of paraquat on mitochondria of rat skeletal muscle. Comp. Biochem. Physiol. 86, 375–378. Yamashita, M., Naito, H., and Takagi, S. (1987). The effectiveness of a cation resin (kayexalate) as an adsorbent of paraquat: experimental and clinical studies. Human Toxicol. 6, 89–90. Yamashita, M., Yamashita, M., and Ando, Y. (2000). A long-term followup of lung function in survivors of paraquat poisoning. Hum. Exp. Toxicol. 19, 99–103. Yang, J. O., Gil, H. W., Kang, M. S., Lee, E. Y., and Hong, S. Y. (2009). Serum total antioxidant statuses of survivors and nonsurvivors after acute paraquat poisoning. Clin. Toxicol. (Phila) 47, 226–229. Yasaka, T., Ohya, I., Matsumoto, J., Shiramizu, T., and Sasaguri, J. (1981). Acceleration of lipid peroxidation in human paraquat poisoning. Arch. Intern. Med. 141, 1169–1171. Yasaka, T., Okudaira, K., Fujito, H., and Shiramitsu, T. (1986). Further studies of lipid-peroxidation in human paraquat poisoning. Arch. Intern. Med. 146, 681–685. Yonemitsu, K. (1986). Pharmacokinetic profile of paraquat following intravenous administration to the rabbit. Forensic. Sci. Int. 32, 33–42. Yonei, S., Noda, A., Tachibana, A., and Akasaka, S. (1986). Mutagenic and cytotoxic effects of oxygen free radicals generated by methyl viologen (paraquat) on Escherichia Coli with different DNA-repair capacities. Mutat. Res. 163, 15–22. Yoshimura, Y., Watanabe, Y., and Shibuya, Y. (1993). Inhibitory effects of calcium channel antagonists on motor dysfunction induced by intracerebroventricular administration of paraquat. Pharmacol. Toxicol. 72, 229–235. Younes, M., Cornelius, S., and Seigers, C. P. (1985). Iron-supported in vivo lipid peroxidation induced by compounds undergoing redox cycling. Chem-Biol. Interact. 54, 97–103. Yu, H. Y., Lai, Y. R., Kuo, T. L., and Shen, Y. Z. (1994). Effects of ethanol on pharmacokinetics and intestinal absorption of paraquat in animals. J. Toxicol. Sci. 19, 67–75. Zavala, D. C., and Rhodes, M. L. (1978). An effect of paraquat on the lungs of rabbits: Its implications in smoking contaminated marijuana. Chest 74, 418–420. Zayed, J., Ducic, S., Campanella, G., Panisset, J. C., Andre, P., Masson, H., and Roy, M. (1990). Facteurs environnementeaux dans l’etiologie de la maladie de Parkinson. Can. J. Neurol. Sci. 17, 286–291. Zhang, X., Cherrington, N. J., and Wright, S. H. (2007). Molecular identification and functional characterization of rabbit MATE1 and MATE2-K. Am. J. Physiol. Renal. Physiol. 293, F360–F370. Zilker, T., Fogt, F., and von Clarmann, M. (1988). Kein Parkinsonsyndrom nach akuter paraquat intoxikation. Klin. Wochenschr. 66, 1138–1141.

Chapter 84

Phenoxy Herbicides (2,4-D) Elke Kennepohl1, Ian C. Munro2 and James S. Bus3 1

Kennepohl Consulting Cantox Health Sciences International 3 The Dow Chemical Company 2

84.1  Introduction Phenoxy herbicides have been commercially available for over 60 years and are the most widely used family of herbicides worldwide. 2,4-Dichlorophenoxyacetic acid (2,4-D), the most common of the phenoxy herbicides, is one of the best-studied agricultural chemicals. This chapter focuses primarily on 2,4-D since it is the most widely used herbicide and the majority of the literature on phenoxy herbicides pertains to studies with 2,4-D. The safety of using phenoxy herbicides was first questioned when a series of case-control studies was published by Lennart Hardell in the late 1970s, in which he hypothesized that the occurrence of three rare forms of cancer (Hodgkin’s disease, soft tissue sarcoma, and non-Hodgkin’s lymphoma) in workers was related to exposure to these herbicides along with dioxins known to contaminate 2,4,5trichlorophenoxyacetic acid (2,4,5-T). Since that time, several human and animal studies have been conducted which do not lend support to his hypothesis. As well, several expert panels have been convened to assess the safety of 2,4-D, and all have concluded that there is no evidence to suggest that 2,4-D poses any risk to human health under its intended conditions of use. In fact, 2,4-D has been classified by the U.S. Environmental Protection Agency (EPA) as a Group D (not classifiable as to human carcinogenicity) because “the evidence is inadequate and cannot be interpreted as showing either the presence or absence of a carcinogenic effect.” Because of the vast amount of data available on 2,4-D, this chapter provides a brief summary and overview of the available studies.

84.2  Physical and chemical properties Several phenoxy acids have been used as herbicides, including 2,4,5-T, 4-(2,4-dichlorophenoxy) butyric acid (2,4-DB), 2-(2,4-dichlorophenoxy propionic acid) (dichlorprop), 2(2-methyl-4-chlorophenoxy) propionic acid (MCPP or Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

mecoprop), 2-methyl-4-chlorophenoxyacetic acid (MCPA), and 2-(2,4,5-trichlorophenoxy) propionic acid (Silvex), with the most commonly and widely used herbicide being 2,4-D. 2,4,5-T and Silvex are no longer manufactured or sold. Figure 84.1 shows the chemical structures of the phenoxy acids.

84.2.1  2,4-D Acid, Salts, and Esters The basic form of 2,4-D is the acid, but 2,4-D is often formulated as an inorganic salt, amine, or ester through various manufacturing processes, and is used in many commercial products. CAS number:

94-75-7

Chemical name:

2,4-dichlorophenoxyacetic acid

Trade names:

AC Aquacide, Acme, Agsco, Alco, Alligare, Alphaset, Aqua-Kleen, Arc-Camba, Bakker, Banvel, Barrage, Bonide, Brash, Butoxone, Butyrac, Candor, Chaser, Chemsico, CIL Lawn, Clean Crop, Confront, Cornbelt, Crossbow, Curtail, Cutback, Dexol, Dissolve, Doom, Double Kill, Double Up, Drexel, Dupon Cimarron Max, Dyvel, Escalade, Esteron, Exit, Fertilizer Plus, Five-Star, Focus, Forefront, Formula 40F, Gharda, Gladeamine, Glymix, Grazon, Green Thumb, Greenleaf, Greensweep, Growell, Hellion, Hilco-X, Hivol, Hoelon, Illoxan, Killex, Knockout, Landmaster, Later’s Weed Stop, Lawn Pro, Lazer, Maestro, Mecoturf, Millenium, Miraclo-Gro, Misty Repco Kill, Momentum, Nufarm, Oasis, Outlaw, Paramount, Pasture, Penoxsulam, Picloram, Proturf, Quadmec, Quincept, Range Star, Rangeland, Real-Kill, Recoil, Restore, Rifle-D, Riverdale, Rustler, Saber, Salvo, Savage, Savana, Selective, Shotgun, Solution, Speed Zone, Spoiler, Starane, Statesman, Strike 3, Super Plus, Target, Thundermaster, Top Gun, Toram, Tordon, Trillion, Trimec, Trooper, Turf Builder, Turf Pride, Turflon, Viper, Viterra, Weco Max, Weed Rhap, Weedar, Weedaway, Weedaxe, Weed-B-Gon, Weedex, Weedhawk, Weedmaster, Weedone, Weedstroy, Winterizer

1829

Hayes’ Handbook of Pesticide Toxicology

1830

CH3 Cl

O

CH2

COOH

Cl

O

Cl

(CH2)3

COOH

Cl

Cl

OCH2COOH

HN

Cl

O

Cl

Cl

O Cl

O

COOH

Cl

O

CH2

C

Cl

CH2 CH3 O

CH2

CH2 H10

CH3

O

CH

COOH

Cl

O

O

CH2

COOH

Cl

Cl 2,4-D 2-ethylhexyl ester

COOH

2,4,5-T

CH3

Cl

CH2

Cl

Mecoprop

Cl

Cl

CH

CH3 2,4-D dimethylamine salt

COOH

Dichlorprop

CH3 CH3 CH3

CH

Cl 2,4-DB

2,4-D acid

Cl

O

MCPA

Silvex

CH2COOCH2CH2OCH2CH2CH2CH2

Cl 2,4-D butoxyethanol ester

Figure 84.1  Chemical structures of phenoxy acid herbicides.

Appearance:

white powder

Molecular weight:

255.49

Empirical formula:

C8H6Cl2O3

Melting point:

156.6°C (pure acid); 150–151°C (technical acid)

Molecular weight:

221.04

Boiling point:

Decomposes

Water solubility:

278 mg/liter at 25°C (acid)

Melting point:

140.5°C

Vapor pressure:

0.022 mm Hg at 25°C

Boiling point:

130°C at 1 mm Hg (isopropyl ester)

Water solubility:

900 mg/liter at 25°C (acid)

Vapor pressure:

0.02 mPa at 25°C (acid)

Partition coefficient:

2.81

84.2.2  2,4,5-T

84.2.3  2,4-DB CAS number:

94-82-6

Chemical name:

4-(2,4-dichlorophenoxy) butyric acid

Trade names:

Butoxone, Butyrac, Butirex, Embutone, Embutox, Legumex, Venceweed

Appearance:

colorless to white crystals

Empirical formula:

C10H10Cl2O3

Molecular weight:

249.10

CAS number:

93–76–5

Chemical name:

2,4,5-trichlorophenoxyacetic acid

Trade names:

no longer manufactured or sold

Melting point:

117–119°C

Appearance:

White crystals

Water solubility:

46 mg/liter at 25°C

Empirical formula:

C8H5Cl3O3

Vapor pressure:

negligible (acid and salts)

Chapter | 84  Phenoxy Herbicides (2,4-D)

84.2.4  Dichlorprop (2,4-DP)

1831

84.2.7  Silvex

CAS number:

120–36–5

CAS number:

93-72-1

Chemical name:

2-(2,4-dichlorophenoxy) propionic acid

Chemical name:

Trade names:

Cornox, Hedonal, Weedone, Estaprop

2-(2,4,5-trichlorophenoxy) propionic acid

Appearance:

colorless crystals

Trade names:

no longer manufactured or sold

Empirical formula:

C9H8Cl2O3

Appearance:

white powder

Molecular weight:

235.07

Empirical formula:

C9H7Cl3O3

Melting point:

117–118°C

Molecular weight:

269.51

Water solubility:

350 mg/liter at 20°C

Melting point:

181.6°C

Water solubility:

200 mg/liter at 25°C

84.2.5  Mecoprop (MCPP) CAS number:

7085-19-0

Chemical name:

2-(4-chloro-2-methylphenoxy) propionic acid

Trade names:

Kilprop, Mecopar, Triester-II, MecAmineD, Triamine-II, Triplet, TriPower, Trimec, Trimec-Encore, U46 KV Fluid

Appearance:

White to light brown crystalline solid

Empirical formula:

C10H11ClO3

Molecular weight:

214.65

Melting point:

93–95°C

Water solubility:

very soluble at 25°C

Vapor pressure:

0.31 mPa at 20°C

Partition coefficient:

1.26 at pH7

84.2.6  Mcpa CAS number:

94-74-6

Chemical name:

2-methyl-4-chlorophenoxyacetic acid

Trade names:

Agritox, Agroxone, Agrozone, Agsco MXL, Banlene, Blesal MC, Bordermaster, Cambilene, Cheyenne, Chimac Oxy, Chiptox, Class MCPA, Cornox Plus, Dakota, Ded-Weed, Empal, Envoy, Legumex, Malerbane, Mayclene, Mephanac, Midox, Phenoxylene, Rhomene, Rhonox, Sanaphen-M, Shamrox, Selectyl, Tiller, Vacate, WeedRhap, Zhelan

Appearance:

colorless crystals

Empirical formula:

C9H9ClO3

Molecular weight:

200.62

Melting point:

118–119°C

Water solubility:

825 mg/liter at 25°C (acid)

Vapor pressure:

0.2 mPa at 20°C

84.3  History of use For over 60 years, 2,4-D has been the most commonly and widely used herbicide throughout the world. When applied to plants, 2,4-D is absorbed through the roots and leaves within 4 to 6 hours and is distributed in the plant via the phloem (WHO, 1984). Once absorbed, 2,4-D selectively eliminates broadleaf plants (due to their larger leaf area and hence, greater absorption) by mimicking the effect of auxins (i.e., plant growth-regulating hormones) and stimulating growth, rejuvenating old cells, and overstimulating young cells, leading to an abnormal growth pattern and death in some plants (Mullison, 1987). In addition, 2,4-D affects plant metabolism, which leads to interference with food transport (Mullison, 1987). 2,4-D is primarily used as a herbicide in agriculture, forestry, and lawn care practices, with the majority (60%) of the total usage in the United States being reported for use as weed control in agriculture (i.e., corn and small grains) (EPA, 1997). 2,4-D is reported to be effective against dandelion, plantain, chickweed, henbit, white clover, heal-all, red sorrel, curly dock, chicory, yellow rocket, speedwell, ground ivy, spurge, oxalis, knotweed, purslane, thistle, wild violet, wild onion, wild garlic, lespedeza, yellow nutsedge, crabgrass, sumac, willow, sagebrush, ragweed, Eurasian water milfoil, and water hyacinth (Lefton et al., 1991; Mullison, 1987; WHO, 1975). To a lesser extent, 2,4-D is used as a growth regulator on various crops ranging from potatoes to citrus fruits (WHO, 1975, 1984). In the past, 2,4-D was combined with 2,4,5-trichlorophenoxyacetic acid (2,4,5-T) for brush and weed control. Over 10 million gallons of a special concentrated mixture called Agent Orange were applied in the Vietnam War to defoliate trees (Wolfe, 1983). A reregistration eligibility review/re-evaluation of 2,4-D was conducted by the U.S. EPA in 2005, and by Health Canada in 2008, with both agencies deciding that 2,4-D was eligible for continued use.

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84.4  Formulations 2,4-D is formulated into end-use products to facilitate application. Water-soluble salts and amines are usually prepared as aqueous solutions with small amounts of additives such as water conditioners and antifoam agents. The oil-soluble esters are often formulated with petroleum solvents (e.g., kerosene or naphtha) plus emulsifiers. Such formulations are then diluted with relatively large amounts of water to make the final herbicide spray mixture. Several forms of 2,4-D can be combined with dry fertilizer ingredients to form lawn “weed and feed” products. In the past, concerns arose regarding possible contamination of 2,4-D formulations with dioxins, notably the polychlorinated dibenzo-p-dioxins (PCDDs) and more specifically 2,3,7,8-tetrachlorodibenzo-p-dioxin (T4CDD), and nitrosamines. 2,4-D formulations have been reported to contain 2,3,7,8-T4CDD but only when 2,4,5-T was present (Cochrane et al., 1981, 1982a, 1982b; Woolson et al., 1972). 2,4-D formulations currently sold in the United States contain very few PCDD contaminants. In fact, analytical studies of the more recent formulations have repeatedly shown dioxin levels to be below the limit of quantitation set by the U.S. Environmental Protection Agency (Berry, 1989; Cramer, 1996). In Canada, a limit of 10 ppb per isomer PCDD (nondetectable for 2,3,7,8-T4CDD at 1 ppb) in 2,4-D formulations has been set (Agriculture Canada, 1983). In an older analysis of 200 samples of various forms of 2,4-D (Cochrane et al., 1981, 1982a, 1982b), all but a few samples tested below the 10-ppb limit. In the past, there was some possible contamination with nitrosamines formed from nitrates used in preserving metal storage containers; however, plastic or epoxy-lining has replaced the metal used for storage containers and nitrosamine formation no longer presents a concern.

Hayes’ Handbook of Pesticide Toxicology

the highest exposures were obtained in occupational settings, with reported average estimated internal doses ranging from 0.01 to 40 g/kg body weight/day for forestry workers, and 0.35 to 6.3 g/kg body weight/day for commercial applicators and farmers (Frank et al., 1985; Grover et al., 1986; Lavy and Mattice, 1984; Lavy et al., 1987; Yeary, 1986). Average estimated internal doses reported to be reached by bystanders or home and garden users were below 0.2 g/kg body weight/day (Harris et al., 1992; Lavy and Mattice, 1984). Most of the past epidemiological studies do not reflect the growing trend toward using protective apparel when applying herbicides. With an increased awareness of worker safety and the new proposed labeling directions, workers are required to wear protective clothing consisting of eye protection, chemical-resistant gloves, long-sleeved shirt, long pants, socks, and shoes. In addition, following use of 2,4-D, it is recommended that workers thoroughly wash their hands, face, and arms with soap and water and wash any contaminated clothing separately. Human urinary biomonitoring studies estimated the absor­ bed dose of 2,4-D in farm applicators as 2.5 to 2.7 g/kg/day (Alexander et al., 2007; Thomas et al., 2009), and in farm family spouses and childen at 0.8 and 0.22 g/kg/day, respectively (Alexander et al., 2007). In a biomonitoring study of 135 preschool-aged children and their adult caregivers, Morgan et al. (2008) estimated the maximum absorbed dose in the chlidren as 0.28 g/kg/day, which, using a Biomonitoring Equivalents exposure analysis, were concluded as well below exposure guidance values developed by the US Environmental Protection Agency (Aylward et al., 2009).

84.6  Toxicological studies 84.6.1  Absorption

84.5  Human exposure to 2,4-D 2,4-D is one of the most commonly used herbicides both domestically and commercially, and exposure to it can occur via inhalation, ingestion, and dermal contact. Respiratory exposure to 2,4-D is less than 2% of total exposure (Grover et al., 1986), and residual levels in foodstuffs or drinking water are essentially nondetectable or only detected in trace amounts (Duggan and Corneliussen, 1972; Duggan and Lipscomb, 1969; Gartrell et al., 1985). By far, dermal contact during use of the product accounts for the greatest potential for exposure, with estimates that approximately 90% of total exposure occurs through dermal exposure (Feldman and Maibach, 1974). Since 2,4-D use is typically seasonal and short-term, the duration of exposure is considered repeated subchronic. In a 1992 review of 2,4-D, Munro et al. (1992) summarized exposures to 2,4-D in a variety of occupational and home-use settings. Based on several epidemiological studies,

2,4-D is rapidly absorbed through the gastrointestinal tract following oral exposure, with peak plasma levels being reached in as little as 10 minutes or up to 24 hours depending on the dose and chemical form of 2,4-D (Erne, 1966a; Khanna and Fang, 1966; Knopp and Schiller, 1992; Kohli et al., 1974; Pelletier et al., 1989; Sauerhoff et al., 1977). The rate of absorption is related to dose, with absorption occurring more rapidly at lower doses (i.e., 0.4 mg/kg body weight/day) than at higher doses (i.e., 1 mg/kg body weight/day) (Pelletier et al., 1989). Absorption of 2,4-D esters has been reported to occur more slowly than for acid or salt forms (Erne, 1966a); however, the excretion rates for the various forms are reported to be similar (Khanna and Fang, 1966; Knopp and Schiller, 1992; Pelletier et al. 1989). Dermal contact is the major route of exposure to 2,4-D. In occupationally exposed humans, dermal absorption was reported to occur rapidly based on the detection of 2,4-D in urine within 4 hours (Feldman and Maibach, 1974), and although the percentage absorbed is variable, it is usually

Chapter | 84  Phenoxy Herbicides (2,4-D)

less than 6% (EPA, 1996; Feldman and Maibach, 1974; Harris and Solomon, 1992). Studies in rats and monkeys showed these percentages to be highly variable and dependent on chemical form, vehicle, and animal species (Grisson et al., 1987; Knopp and Schiller, 1992; Moody et al., 1990, 1991; Pelletier et al., 1989, 1990). Although no controlled studies have been conducted to assess the absorption rate via inhalation exposure, epidemiological studies of occupationally exposed workers indicate that absorption is rapid by both dermal and inhalation routes (Frank et al., 1985; Kolmodin-Hedman and Erne, 1980).

84.6.2  Distribution 2,4-D is highly water soluble and therefore is widely distributed, but does not accumulate, in the body. It also does not readily cross lipid membranes, and at physiological pH, it exists predominately in the ionized form. 2,4-D uses active transport systems to enter tissues and cross the blood/ brain barrier (Kim and O’Tuama, 1981; Pritchard, 1980). Another factor which contributes to the extent of tissue distribution of 2,4-D is its ability to bind to serum proteins (Erne, 1966a; Fang and Lindstrom, 1980; Orberg, 1980). Peak tissue levels in rats have been reported anywhere from 10 minutes to 8 hours depending on the dose administered (0.4 to 240 mg/kg body weight) (Khanna and Fang, 1966; Pelletier et al., 1989). Following exposure, 2,4-D has been detected in liver, kidney, and lung of a variety of animal species (Clark et al., 1975; Erne, 1966a). Levels in brain were reported to account for only a very small percentage of the exposure dose (Erne, 1966a; Tyynela et al., 1990); however, at levels of intoxication (i.e., 300 mg/kg body weight, which is well above the level of renal saturation), levels in brain and cerebrospinal fluid of rats were increased relative to plasma levels (Elo and Ylitalo, 1977, 1979; Tyynela et al., 1990). At these high dose levels, the organic acid transport system responsible for the efflux of 2,4-D out of the brain is inhibited (Kim et al., 1983; Pritchard, 1980; Tyynela et al., 1990; Ylitalo et al., 1990). In addition, vascular damage has been reported in rats administered extremely high doses of 2,4-D (i.e., more than 300 mg/kg body weight) (Elo et al., 1988), which may facilitate an increased influx of 2,4-D through the compromised blood/brain barrier (Elo et al., 1988; Hervonen et al., 1982; Tyynela et al., 1990). Saturation of plasma protein binding also may contribute slightly to the increased brain: blood ratio of 2,4-D reported in rats at these exposure levels (Tyynela et al., 1990; Ylitalo et al., 1990). 2,4-D also has been reported to pass the placental barrier in mice, rats, and pigs, and has been detected in the uterus, placenta, fetus, and intrauterine fluid of exposed animals (Erne, 1966a; Fedorova and Belova, 1974; Lindquist and Ullberg, 1971) but was rapidly eliminated (Lindquist and Ullberg, 1971).

1833

84.6.3  Pharmacokinetics Depending on the chemical form of 2,4-D and the animal species tested, plasma half-lives following oral exposure of 100 mg/kg body weight range from 3.5 to 12 hours (Erne, 1966a). Lower doses (i.e., 3 mg/kg body weight) in rats showed half-lives of 0.5 to 0.8 hours (Khanna and Fang, 1966), indicating that clearance rates are highly dependent on dose. In human studies, plasma clearance of orally administered 2,4-D was found to follow first-order kinetics with urinary excretion half-lives ranging from 10.2 to 28.4 hours (Sauerhoff et al., 1977), which is consistent with the findings (urinary excretion half-life  18 hours) from a forestry worker who exhibited the highest amount of 2,4-D excretion among two groups exposed over a period of 11 or 18 days (Frank et al., 1985). The pharmacokinetics of 2,4-D following dermal absorption is apparently different from that following the oral route (Pelletier et al., 1989). Plasma levels tend to reach a plateau and decline more rapidly following oral exposure. In addition, plasma clearance has been reported to follow biphasic kinetics beginning 8 hours post-dosing, with half-lives for various tissues ranging from 0.6 to 2.3 hours for the first phase and 25.7 to 29 hours for the second phase. Furthermore, cumulative urinary excretion of 2,4-D increases more slowly following dermal rather than oral exposure.

84.6.4  Metabolism Once absorbed into body fluids and tissues, the salts and esters of 2,4-D undergo acid and/or enzymatic hydrolyzation to form 2,4-D acid. In laboratory animals and humans following oral exposures, the presence of acidhydrolyzable conjugates has been reported at 0 to 27% of the administered 2,4-D (Erne, 1966b; Grunow and Bohme, 1974; Kohli et al., 1974; Sauerhoff et al., 1977). The available data indicate that 2,4-D is not metabolized to reactive intermediates and is excreted predominately as the parent compound.

84.6.5  Excretion Regardless of the route of exposure, 2,4-D is predominately excreted in the urine (Erne, 1966a; Feldman and Maibach, 1974; Khanna and Fang, 1966; Knopp and Schiller, 1992; Moody et al., 1990, 1991; Pelletier et al., 1989). The rate of urinary excretion is inversely proportional to dose. For example, at oral doses of 3 to 30 mg/kg body weight given to rats, 93 to 96% of the dose was excreted within 48 hours, whereas at higher doses (i.e., 60 mg/kg body weight) the percentage of dose excreted within 24 hours decreased linearly with increasing dose (Khanna and Fang, 1966). In rats, 90% of oral doses of 30 mg/kg body weight or less were excreted in the urine within 24 hours (Khanna

Hayes’ Handbook of Pesticide Toxicology

1834

and Fang, 1966; Knopp and Schiller, 1992; Pelletier et al., 1989). Similarly, in humans administered an oral dose of 5 mg 2,4-D/kg body weight, 77% of the dose was excreted within 96 hours (Kohli et al., 1974) and 87 to 100% of the dose was excreted in the urine over 6 days (Sauerhoff et al., 1977). 2,4-D is predominately excreted by the kidney using an active transport system. Saturation of renal clearance appears to occur at 50 to 60 mg/kg body weight (Gorzinski et al., 1987; Khanna and Fang, 1966) based on kidney concentrations and urinary excretion rates. Another significant route of excretion in occupationally exposed workers is perspiration (Sell et al., 1982). Following a 2-hour exposure, 2,4-D was detected in T-shirt extracts (i.e., measure of perspiration) for 2 weeks and in urine for 5 days. 2,4-D also has been reported to be excreted in the milk of lactating rats exposed to 2,4-D (Fedorova and Belova, 1974).

84.6.6 Animal Studies 84.6.6.1  Acute Toxicity Numerous acute toxicological tests have been conducted on the various forms of 2,4-D as summarized in Table 84.1. Overall, oral exposure to 2,4-D shows moderate to low toxicity, whereas dermal and inhalation toxicity are low. Dermal irritation in rabbits was considered slight for the acid form of 2,4-D and minimal for the salt and ester forms. Reported eye irritation in rabbits, on the other hand, is severe for the acid and salt forms, but minimal for the ester.

84.6.6.2  Subchronic Toxicity Further to the acute toxicity data, numerous subchronic studies have been conducted on a variety of 2,4-D forms, by different exposure routes, and in various animal species. The subchronic studies conducted range from 3-week dermal studies in rabbits to 13-week dietary studies in dogs and rodents. The results of these studies are summarized in Table 84.2. Overall, at doses above the threshold

of saturation for renal clearance, the key target organs in rats appear to be primarily the kidney and, to some extent, the thyroid. The changes reported in the rat kidney were the loss of epithelial cells in the proximal tubule brush border; the changes in the thyroid were follicular cell hypertrophy in association with a reduction in serum thyroxine levels. These changes were consistent over all forms of 2,4-D tested, with a reported no-observed-adverse-effect level (NOAEL) of 15 mg/kg body weight/day (Charles et al., 1996b; Szabo and Rachunek, 1991; Yano et al., 1991a, 1991b). Some of these findings were reported at lower doses in older rat studies using the acid form of 2,4-D (Gorzinski et al., 1981a, 1981b); however, the histological effects were not statistically significant at 15 mg/kg body weight/day, which is consistent with the more recent studies. In another older study (Serota, 1983a), other minor histological effects in the kidney were reported in male rats at 5 mg/kg body weight/day and in female rats at 1 mg/kg body weight/day. These effects were not reported in any other of the subchronic studies. Although some thyroid changes have been reported at 15 mg/kg body weight/day in rats (Serota, 1983a), these changes were considered incidental and do not affect the conclusion that the subchronic NOAEL for 2,4-D is 15 mg/kg body weight/day (Munro et al., 1992). Similar results with respect to the kidney have been reported in 13-week dietary mouse studies (Serota, 1983b; Schulze, 1991). In a 13-week dog study in which 2,4-D was administered by gelatin capsules, kidney effects consisting of reduced cytoplasmic eosinophilia of the epithelial cells lining some convoluted tubules were reported at lower doses, resulting in a NOAEL of 1 mg/kg body weight/day (ITF, 1990). Thyroid changes were not reported in the mouse or dog.

84.6.6.3  Reproductive and Developmental Toxicity Several multigenerational and developmental animal studies have been conducted to assess the potential of 2,4-D to affect reproduction and the developing fetus, and are summarized in

Table 84.1  Acute Toxicitya Involving Various Chemical Forms of 2,4-D Oral LD50 (mg/kg bw/d)

Dermal LD50 (mg/ kg bw/d)

Inhalation LD50 (mg/liter)

Form of 2,4-D

Rat

Mouse

Dog

Guinea pig

Chicken

Rat

Rabbit

Rat

Acid

na

1400–  2000

1.79

639–980

312–434

25–250

397–553

358–817

Salt

863–  2000

na

na

na

na

2000

2000

3.8

Ester

440–982

na

na

na

na

na

1829–  2000

4.6

bw  body weight. na  not available. a Condensed from Munro et al. (1992).

Chapter | 84  Phenoxy Herbicides (2,4-D)

1835

Table 84.2  Summarya of Subchronic Studies on 2,4-D Tested in Various Animal Species Route

Dose (mg/kg bw/d) [in acid equivalents]

NOAELb (mg/kg bw/d)

Reference

2,4-D (100% pure)

diet

0, 15, 60, 100, 150

15 (females only)

Gorzinski et al. (1981a)

2,4-D (97.5% pure)

diet

0, 15, 60, 100, 150

15 (females only)

Gorzinski et al. (1981b)

2,4-D (97.5% pure)

diet

0, 1, 5, 15, 45

1 (males only)

Serota (1983a)

2,4-D (96.1% pure)

diet

0, 1, 15, 100, 300

15

Charles et al. (1996b)

2,4-D ethylhexyl ester

diet

0, 1, 15, 100, 300

15

Charles et al. (1996b)

2,4-D dimethylamine salt

diet

0, 1, 15, 100, 300

15

Charles et al. (1996b) Szabo and Rachunek (1991)

Chemical species 13-week: Rat

2,4-D butoxyethyl ester

diet

0, 1, 15, 100, 300

15

2,4-D triisopropanolamine salt

diet

0, 1, 15, 100, 300

15

Yano et al. (1991b)

2,4-D isopropylamine salt

diet

0, 1, 15, 100, 300

15

Yano et al. (1991a)

intraperitoneal

0, 100, 150

—

Lukowicz-Ratajczak and Krechniak (1988)

2,4-D acid

diet

0, 5, 15, 45, 90

—

Serota (1983b)

2,4-D acid

diet

0, 1, 15, 100, 300

15

Schulze (1991)

oral intubation

50, 100, 200

—

Kuntz et al. (1990)

diet

100

—

Lundgren et al. (1987)

2,4-D acid

gelatin capsule

0, 0.3, 1, 3, 10

1

ITF (1990)

2,4-D acid

diet

0, 0.5, 1.0, 3.75, 7.5

1

Charles et al. (1996c)

2,4-D dimethylamine salt

diet

0, 1.0, 3.75, 7.5

1

Charles et al. (1996c)

2,4-D 2-ethylhexyl ester

diet

0, 1.0, 3.75, 7.5

1

Charles et al. (1996c)

2,4-D acid

dermal

0, 10, 100, 1000

1000

Schulze (1990a)

2,4-D dimethylamine salt

dermal

0, 10, 100, 300

10

Schulze (1990b)

2,4-D 2-ethylhexyl ester

dermal

0, 10, 100, 1000

10

Schulze (1990c)

2,4-D triisopropanolamine salt

dermal

0, 55, 193, 553

553

Mizell et al. (1990b)

2,4-D isopropylamine salt

dermal

0, 39, 98, 275

275

Mizell et al. (1990a)

2,4-D butoxyethyl ester

dermal

0, 32, 96, 321

321

Mizell et al. (1989)

12-week: Rat 2,4-D sodium salt 13-week: Mouse

14-day: Mouse 2,4-D acid 4-day: Mouse 2,4-D acid 13-week: Dog

21-day: Rabbit

a

Adapted from Munro et al. (1992). NOAEL  no-observed-adverse-effect level.

b

Table 84.3. In general, the results of the available studies indicate that 2,4-D is not teratogenic and does not affect reproduction except at maternally toxic doses or those saturating the threshold for renal clearance (i.e., 50 mg/kg body weight/day). At doses above the maximum tolerated dose (MTD), some

developmental effects have been reported in test animals (i.e., decreased fetal weight gain, increased incidence of lumbar ribs and wavy ribs, and delayed ossification of bone). The only teratogenic effect (i.e., cleft palate) reported was in mice, but occurred only at maternally toxic doses.

Hayes’ Handbook of Pesticide Toxicology

1836

Table 84.3  Summary of Developmental and Reproductive Toxicity Studies on 2,4-D Tested in Various Animal Species Route

Exposure duration

Dose (mg/kg bw/d)

NOAELa (mg/kg bw/d)

Reference

2,4-D triisopropanolamine salt

gavage

GDb 7–19

0, 10, 30, or 75c

10d; 75e

Liberacki et al. (1994), Charles et al. (1996a)

2,4-D isopropylamine salt

gavage

GD 7–19

0, 10, 30, or 75c

10d; 75e

Liberacki et al. (1994), Charles et al. (1996a)

2,4-D butoxyethyl ester

gavage

GD 7–19

0, 10, 30, or 75c

10d; 75e

Liberacki et al. (1994), Charles et al. (1996a)

2,4-D acid

gavage

GD 6–18

0, 10, 30, or 90

30d; 90e

Hoberman (1990), Charles et al. (1996a)

2,4-D dimethylamine salt

gavage

GD 6–18

0, 10, 30, or 90c

30d; 90e

Martin (1991), Charles et al. (1996a)

2,4-D ethylhexyl ester

gavage

GD 6–18

0, 10, 30, or 75c

30d; 75e

Martin (1992a), Charles et al. (1996a)

2,4-D diethanolamine

gavage

GD 6–19

0, 10.2, 20.3, or 40.6c

10.2d; 40.6e

2,4-D isooctyl ester

oral

GD 6–15

0, 12.5, 25, 50, 75, or 87.5c

87.5d; 25e

Schwetz et al. (1971)

2,4-D propylene glycol butyl ether

oral

GD 6–15

0, 12.5, 25, 50, 75, or 87.5c

87.5d; 25e

Schwetz et al. (1971)

2,4-D acid

oral

GD 6–15

0, 12.5, 25, 50, 75, or 87.5

87.5d; 25­

Schwetz et al. (1971)

2,4-D isooctyl ester

oral

GD 5–8

0, or 87.5c

87.5d,e

Schwetz et al. (1971)

2,4-D propylene glycol butyl ether

oral

GD 5–8

0, or 87.5c

87.5d,e

Schwetz et al. (1971)

2,4-D isooctyl ester

oral

GD 8–11

0, 50, or 87.5c

87.5d; 50e

Schwetz et al. (1971)

2,4-D isooctyl ester

oral

GD 12–15

0, 50, or 87.5c

87.5d,e

Schwetz et al. (1971)

2,4-D isooctyl ester

oral

GD 6–15

0, 50, or 150

150d; 50e

Khera and McKinley (1972)

2,4-D butyl ester

oral

GD 6–15

0, 50, or 150

150d; 50e

Khera and McKinley (1972)

2,4-D butoxyethynol

oral

GD 6–15

0, 50, or 150

150d; 50e

Khera and McKinley (1972)

2,4-D dimethylamine salt (49.5%)

oral

GD 6–15

0, 100, or 300

300d; 50e

Khera and McKinley (1972)

2,4-D acid

oral

GD 6–15

0, 50, or 100

100d; 50e

Khera and McKinley (1972)

2,4-D acid

oral

GD 6–15

0, 25, 50, or 100

100d; 50e

Khera and McKinley (1972)

2,4-D acid

oral

GD 6–15

0, 25, 50, 100, or 150

150d; 50e

Khera and McKinley (1972)

Chemical species Developmental: Rabbit

Developmental: Rat

Chapter | 84  Phenoxy Herbicides (2,4-D)

1837

Table 84.3  (Continued) 2,4-D propylene glycol butyl ether

oral

GD 6–15

0, 6.25, 12.5, 25, or 87.5c

87.5d; 25e

Unger et al. (1981)

2,4-D isooctyl ester

oral

GD 6–15

0, 6.25, 12.5, 25, or 87.5c

87.5d; 25e

Unger et al. (1981)

2,4-D acid

gavage

GD 6–15

0, or 115

115d,e

Chernoff et al. (1990)

2,4-D ethylhexyl ester

gavage

GD 6–15

0, 10, 30, or 90c

10d, 30e

Martin (1992b), Charles et al. (1996a)

2,4-D dimethylamine salt

gavage

GD 6–15

0, 12.5, 50, or 100c

12.5d; 50e

Lochry (1990), Charles et al. (1996a)

2,4-D acid

gavage

GD 6–15

0, 8, 25, or 75

75d,e

Nemec et al. (1983), Charles et al. (1996a)

oral

GD 7–15

0, or 0.56 mM/kg bw

0.56d; 0.56e

Developmental: Mouse 2,4-D acid 2,4-D acid

oral

GD 11–14

0, or 0.80 mM/kg bw

0.80

2,4-D acid

oral

GD 12–15

0, or 1 mM/kg bw

1d,e d,e

d,e

subcutaneous

GD 12–15

0, or 1 mM/kg bw

2,4-D isopropyl ester

oral

GD 7–15

0, or 0.56 mM/kg bw

0.56

d,e

2,4-D n-butyl ester

oral

GD 7–15

0, or 0.56 mM/kg bw

2,4-D n-butyl ester

oral

GD 12–15

0, or 1 mM/kg bw

2,4-D isooctyl ester

oral

GD 7–15

0, or 0.56 mM/kg bw

0.56 1

Courtney (1977) Courtney (1977)

2,4-D acid

1

Courtney (1977)

Courtney (1977) d,e

d,e

Courtney (1977) Courtney (1977) Courtney (1977)

0.56

d,e

Courtney (1977)

d,e

Courtney (1977)

2,4-D propylene glycol butyl ether

oral

GD 7–15

0, or 0.56 mM/kg bw

0.56

2,4-D propylene glycol butyl ether

oral

GD 12–15

0, or 1 mM/kg bw

1d; 1e

Courtney (1977)

d,e

Kavlock et al. (1987)

2,4-D acid

oral

GD 8–12

0, or 87.5

87.5

2,4-D propylene glycol butyl ether

oral

GD 8–12

0, or 87.5

87.5d,e

Kavlock et al. (1987)

2,4-D isooctyl ester

oral

GD 8–12

0, or 87.5

87.5d,e

Kavlock et al. (1987)

2,4-D acid

subcutaneous

GD 6–14

0, or 100

100d,e

Bionetics (1968)

2,4-D acid

subcutaneous

GD 6–14

0, or 98

98d,e

Bionetics (1968)

d,e

2,4-D acid

subcutaneous

GD 6–14

0, or 215

215

2,4-D acid

subcutaneous

GD 6–14

0, or 50

50d,e

2,4-D acid

Bionetics (1968) Bionetics (1968)

d,e

oral

GD 6–14

0, or 100

100

Bionetics (1968)

2,4-D

oral

GD 6–10

0, 20, 40, 60, or 100

100e

Collins and Williams (1971)

2,4-D

oral

GD 6–10

0, 40, 60, or 100

100e

Collins and Williams (1971)

2,4-D

oral

GD 6–10

0, 40, 60, or 100

40

Collins and Williams (1971)

diet

2-generation

0, 5, 20, or 80

20d; 5e

Rodwell (1984)

Developmental: Hamster

Reproductive Toxicity: Rat 2,4-D acid a

NOAEL  no-observed-adverse-effect level. b GD  gestational days. c 2,4-D acid equivalents. d Maternal. e Fetal.

1838

The potential testicular and ovarian toxicity of 2,4-D has been extensively evaluated in a series of subchronic and chronic studies in rats. In rats fed 0, 1, 15, 100, or 300 mg/kg body weight/day of either 2,4-D acid, 2,4-D dimethylamine salt, or 2,4-D 2-ethylhexyl ester for 90 days, only minimal effects were noted in testes at the top dose of 300 mg/kg body weight/day (Charles et al., 1996b). These effects, consisting of decreased testes/body weight ratios accompanied by slight histological evidence of atrophy, occurred only at a dose which exceeded the MTD. The NOAEL for testicular effects was 100 mg/kg body weight/day, while the overall NOAEL for the subchronic studies was 15 mg/kg body weight/day based primarily on minor effects in the kidney. No 2,4-D-induced toxicity was reported in the ovaries at any dose. In a subsequent chronic toxicity/oncogenicity study conducted in rats with 2,4-D acid at doses of 0, 5, 75, or 150 mg/kg body weight/day, no treatment-related effects were reported in testes or ovaries at any dose level (Charles et al., 1996a). The findings from these studies were consistent with observations from a series of earlier 90-day (Gorzinski et al., 1987; Serota, 1983a, b; Serrone et al., 1991; Szabo and Rachunek, 1991; Yano et al., 1991a, 1991b) and chronic studies (Charles et al., 1996a; Serota, 1986) conducted with 2,4-D acid and a variety of salt and ester derivatives. The testicular and ovarian toxicity of 2,4-D acid and its dimethylamine salt and 2-ethylhexyl ester also have been examined in a series of subchronic and chronic studies in beagle dogs. Dogs administered either 0, 1, 3.75, or 7.5 mg/kg body weight/day of 2,4-D acid, 2,4-D dimethylamine salt, or 2,4-D 2-ethylhexyl ester for 13 weeks exhibited decreased testes weights at the two highest dose levels (Charles et al., 1996c). The toxicological significance of these findings is uncertain since the organ weight changes were not accompanied by any corroborative histological changes. In addition, both of the two high-dose group animals exhibited body weight gain depressions of approximately 30 to 85%. No 2,4-D-induced effects were seen in the ovaries. A NOAEL of 1.0 mg/kg body weight/ day was established from these studies based on effects in the kidneys. In a follow-up 1-year chronic study conducted at dietary doses of 0, 1, 5, or 7.5 mg/kg body weight/day of 2,4-D acid, no testes or ovary alterations were reported (Charles et al., 1996c). The findings from these dog studies were consistent with earlier 13-week (ITF, 1990) and 2-year dog studies (Hansen et al., 1971). The general lack of 2,4-D-associated testicular toxicity is entirely consistent with the failure of 2,4-D to induce changes in reproductive performance.

84.6.6.4  Immunotoxicity Several subchronic and chronic toxicity studies have provided no evidence from hematological, clinical chemistry, or histopathologic evaluations that 2,4-D is likely to

Hayes’ Handbook of Pesticide Toxicology

induce immune system dysfunction (Charles et al., 1996a, 1996b, 1996c; Szabo and Rachunek, 1991; Yano et al., 1991a, 1991b). Studies that were conducted specifically to examine the possible impact of 2,4-D on various immune system functional parameters have not provided definitive results (Blakley, 1986; Blakley and Blakley, 1986; Blakley and Schiefer, 1986; Zhamsaranova et al., 1987). These studies are difficult to interpret in that the results (1) were inconsistent when evaluated by different routes of exposure, or by comparison of findings from acute and subchronic test regimes; (2) often not reproducible; and (3) not accompanied by adequate descriptions of the normal range of immune parameter values in the test animal populations (Munro et al., 1992). Dermal exposure to 2,4-D acid, salts, or esters also has not been associated with delayed contact hypersensitivity in guinea pigs (Carreon et al., 1983; Carreon and Rao, 1985; Gargus, 1986; Jeffrey, 1986; Jeffrey and Rao, 1986; Schultz et al., 1990).

84.6.6.5  Neurotoxicity Overall, it may be concluded that 2,4-D has little, if any, potential to induce adverse effects in the nervous system at doses that do not cause overt systemic toxicity or that do not saturate processes involved with tissue clearance and renal excretion. No lesions or overt clinical signs of central nervous system toxicity were observed in any of the subchronic toxicity studies in rats (Charles et al., 1996b; Szabo and Rachunek, 1991; Yano et al., 1991a, 1991b) or mice (Schulze, 1991), at doses up to 300 to 560 mg/kg body weight/day. In a chronic rat study designed specifically to investigate the impact of 2,4-D on the nervous system, several neurological parameters were assessed, of which only forelimb grip strength was altered, to a minimal degree, at the highest dose tested, 150 mg/kg body weight/day (Jeffries et al., 1994). Other animal studies have yielded electromyogram results considered indicative of skeletal muscle myotonia following administration of high doses of 2,4-D (50 to 100 mg/kg body weight/day) (Arnold et al., 1991; Beasley et al., 1991; Elo and MacDonald, 1989; Kwiecinski, 1981; Steiss et al., 1987; Toyoshima et al., 1985). In a subchronic toxicity study on 2,4-D 2-ethylhexyl ester, some of the highdose animals were reported to show clinical signs that could possibly be related to myotonia (e.g., hunched posture, languid behavior, ataxia) (Charles et al., 1996b). Myotonia induced by high levels of exposure to 2,4-D does not appear to be the result of toxicological action upon the central nervous system (Buslovich and Pichugin, 1983), but appears to be due to effects mediated at the junction of skeletal muscle nerves and muscle tissue. The biochemical mechanism involved in the induction of myotonia in experimental animals is not well understood; however, according to Rudel and Senges (1972), alteration of chloride ion conductance in muscle fibers appears to be

Chapter | 84  Phenoxy Herbicides (2,4-D)

involved. Because the development of myotonia in animals exposed to high doses of 2,4-D was not accompanied by any pathological effects, and because the reporting of myotonia is restricted to dose levels greater than the threshold for saturation of renal tubular secretion, the effects reported in animal studies are not considered to be indicative of a potential of 2,4-D to induce peripheral polyneuropathy in humans. In an independent review of the literature, it was concluded that exposure to 2,4-D does not produce polyneuropathy in humans nor does polyneuropathy occur in several animal species exposed to high levels of 2,4-D (Mattsson and Eisenbrandt, 1990). Using tests of neurobehavioral parameters (such as the Functional Observational Battery), decreased activity levels, behavioral changes, and motor skill abnormalities have been reported in rats at doses greater than 60 mg/kg body weight/day and in rabbits at doses of approximately 30 mg/kg body weight/day (Breslin et al., 1991; de Duffard et al., 1990b; Duffard et al., 1995; Hoberman, 1990; Jeffries et al., 1994; Liberacki et al., 1991; Martin, 1991; Mattsson et al., 1994, 1997; Oliveira and Palermo-Neto, 1993; Rodwell, 1991; Schulze and Dougherty, 1988; Zablotny et al., 1991). In acute studies conducted using beagle dogs, clinical signs of central nervous system depression and/or abnormalities in electroencephalograms were only reported at doses of 175 mg/kg or greater (Arnold et al., 1991). Rats exposed to high doses of 2,4-D n-butyl ester were reported to display alterations in neurotransmitter concentrations in the brain (de Duffard et al., 1990a; Elo and MacDonald, 1989; Oliveira and Palermo-Neto, 1993). These neurochemical alterations have been hypothesized to result from compromise of the blood-brain barrier by high doses of 2,4-D (Elo et al., 1988; Tyynela et al., 1986). These authors reported that doses of 2,4-D greater than 150 mg/kg body weight resulted in extravasation of albumin in various areas of the brain. Several investigators have reported accumulation of 2,4-D in the brain or cerebrospinal fluid, following administration of high doses (40 to 300 mg/kg body weight) of 2,4-D (Elo and Ylitalo, 1977, 1979; Kim et al., 1988; Oliveira and Palermo-Neto, 1993; Tyynela et al., 1990). Kim et al. (1988) suggested that the increased accumulation of 2,4-D in the brain at high doses was likely not the result of increased permeability of the blood-brain barrier since the entry of the organic solute, 2-deoxyglucose, into rabbit brain was unaffected by 2,4-D pretreatment. Instead, it has been hypothesized that reduced elimination of 2,4-D from the brain via the choroid plexus through competitive inhibition of the organic acid transport pathway was likely responsible for the increased accumulation of 2,4-D (Kim et al., 1988; Ylitalo et al., 1990). The organic acid transport pathway normally actively eliminates acidic metabolites from the brain through the blood-brain barrier. At doses below the capacity of normal renal clearance, there is no evidence in experimental animals to indicate

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that 2,4-D can have an impact on the nervous system. In fact, no clinically observable adverse effects on the nervous system have been observed in animals at doses below 10 to 30 mg/kg body weight, even in long-term studies.

84.6.6.6  Chronic Toxicity and Carcinogenicity Several long-term bioassays have been conducted in rats, mice, and dogs (Arkhipov and Kozlova, 1974; Charles et al., 1996a, 1996c; Innes et al., 1969; Hansen et al., 1971; Serota et al., 1986, 1987). There has been no evidence to suggest that 2,4-D acts as a carcinogen in any of these species. Rats In one older 2-year rat feeding study (Serota et al., 1986), an increase in the incidence of brain astrocytomas was reported in male rats only at the highest dose tested of 45 mg/kg body weight/day; however, the biological characteristics of the tumors were not consistent with chemical carcinogenesis. Moreover, based on the lack of decreased latency, the lack of increased multiplicity, the lack of increased severity, the lack of preneoplastic or target organ effects, the restriction of tumor development to one species and sex, the intergroup variability exhibited among historical controls, the lack of a plausible mechanism of tumorigenesis, the low exposure of the brain to 2,4-D compared to other tissues, and the fact that these tumors have not been reproduced in subsequent studies, it is unlikely that the increased incidence of brain astrocytomas reported by Serota et al. (1986) was related to 2,4-D treatment. In another older study (Hansen et al., 1971) in which rats were fed up to 62.5 mg 2,4-D/kg body weight/day for 2 years, an overall increase in the number of randomly distributed tumors was reported to be statistically significant for male rats. As discussed by Munro et al. (1992), this study was not considered to provide any evidence that 2,4-D is carcinogenic in the rat since it did not meet the requirements of Good Laboratory Practice (GLP) standards, the dose groups were fairly small, the maximum tolerated dose (MTD) was not achieved, and the microscopic examination was not comprehensive. In a third feeding study (Arkhipov and Kozlova, 1974), rats were fed 10% of the reported LD50 (details of dosing not reported) with no significant increase in tumor incidence. More recently, a 2-year GLP-compliant study in which rats were fed 5 to 150 mg 2,4-D/kg body weight/day was completed without any evidence of carcinogenicity (Charles et al., 1996a). In particular, there was no increased incidence of brain astrocytomas even at the MTD. Noncancer endpoints reported in the animals were very similar to those reported in the subchronic studies (Charles et al., 1996b), and a NOAEL of 5 mg/kg body weight/day was established based on increased thyroid weight. Mice No evidence of carcinogenicity has been reported in three long-term mouse studies (Charles et al., 1996a; Innes et al., 1969; Serota et al., 1987). In the first study

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(Innes et al., 1969), mice were orally administered one of three esters of 2,4-D (i.e., isopropyl, butyl, or isooctyl ester) at a dose of 46.4 mg/kg body weight/day for 18 months. No increase in tumor incidence was reported. In the second study (Serota et al., 1987), mice were fed 1 to 45 mg 2,4-D/kg body weight/day for 106 weeks with no evidence of carcinogenicity or other treatment-related effects. In male mice administered the highest two doses (15 and 45 mg/kg body weight/day), an increase in cytoplasmic homogeneity in the renal tubular epithelium was reported. In the third and most recent GLP-compliant study (Charles et al., 1996a), female and male mice were fed 5 to 300 and 5 to 125 mg 2,4-D/kg body weight/day, respectively, for 2 years without any evidence of tumorigenesis. Noncancer effects were limited to slight depression of red blood cell parameters, minor organ weight changes, and histopathological renal effects at the top two doses in both sexes. Dogs Similar to the results from the rodent studies, there has been no evidence to suggest that 2,4-D has carcinogenic potential in dogs (Charles et al., 1996c; Hansen et al., 1971). The results of the long-term studies in dogs support the results reported in subchronic studies. In the older study by Hansen et al. (1971), small groups of beagle dogs were fed 2,4-D in the diet at concentrations reaching 500 ppm over a 2-year period. Following gross and microscopic examinations of several tissues and organs, no lesions related to 2,4-D treatment were reported. The more recent study by Charles et al. (1996c) examined the effects of feeding 0, 1, 5, or 7.5 mg 2,4-D/kg body weight/day to beagle dogs for a period of 52 weeks. The reported effects included body weight gain reduction in females, notably at the highest dose, some serum chemistry alterations (i.e., increased urea nitrogen, creatinine, cholesterol, and alanine aminotransferase activity) in the two highest dose groups, and some histopathological alterations (i.e., perivascular chronic active inflammation of the liver and an increase of pigment in tubular epithelium in both sexes in the two highest dose groups, and pigment in the sinusoidal lining cells in the females of the two highest dose groups). Overall, 2,4-D administration was well tolerated and produced no effects on clinical signs, hematology, urinalysis, or gross necropsy. A NOAEL of 1 mg/kg body weight/day was suggested by the authors.

84.6.7  Genotoxicity Numerous in vitro and in vivo genotoxicity studies have been conducted with 2,4-D. Overall, the results indicate that 2,4-D has very little genotoxic potential. This conclusion has been reached in previous reviews of 2,4-D (CCT, 1987; EPA, 1997; Munro et al., 1992) and is consistent with metabolism studies which have indicated that 2,4-D does not metabolize to reactive intermediates.

Hayes’ Handbook of Pesticide Toxicology

With very few exceptions, bacterial mutagenicity tests using Salmonella typhimurium and Escherichia coli have produced negative results with 2,4-D (Anderson and Styles, 1978; Charles et al., 1996a; Ercegovich and Rashid, 1977; Kappas, 1988; Kappas and Markaki, 1988; Kappas et al., 1984; Mersch-Sundermann et al., 1989; Rashid, 1979; Rashid and Mumma, 1986; Rashid et al., 1984; Simmon et al., 1977; Soler-Niedziela et al., 1988; Styles, 1973; Waters et al., 1980). In yeast cells, mitotic gene conversion and recombination has been reported, but was highly dependent on pH and occurred only at pH 4.3 (Simmon et al., 1977; Waters et al., 1980; Zetterberg, 1978; Zetterberg et al., 1977). Negative or weakly positive results were reported in unscheduled DNA repair and sister chromatid exchange (SCE) assays with mammalian cell systems (Charles et al., 1996a; Clausen et al., 1990; Galloway et al., 1987; Jacobi and Witte, 1991; Styles, 1977; Waters et al., 1980). For the most part, the weakly positive results occurred only in conjunction with cytotoxicity (Clausen et al., 1990; Korte and Jalal, 1982). Similar to the in vitro studies, the majority of in vivo studies with animals, using the most accepted and validated procedures, have produced negative results (Munro et al., 1992). Some studies with occupationally exposed individuals provided marginally positive results in lymphocytes but could not be directly correlated with 2,4-D exposure due to various confounding factors (e.g., age, sex, race, lifestyle habits, etc.) (Crossen et al., 1978; Kaye et al., 1985; Yoder et al., 1973). Several other studies have shown that 2,4-D exposure has no effect on chromosomal aberration or SCE frequency (Charles et al., 1996a, b; Hogstedt and Westerlund, 1980; Linnainmaa, 1983a, 1983b, 1984; Mulcahy, 1980; Mustonen et al., 1986, 1989).

84.7  Studies in humans There has been some concern over a possible association between exposure to 2,4-D and the development of cancer, specifically non-Hodgkin’s lymphoma, Hodgkin’s disease, and soft tissue sarcoma. Comprehensive critical reviews of the epidemiology literature however, have concluded that no causal association between 2,4-D and human cancer has been convincingly documented (Munro et al., 1992; Garabrant and Philbert, 2002). In overview, the majority of the studies were conducted with farmers, forestry workers, and other similar groups of potential users of herbicides. In most of the studies, there were methodological shortcomings in conducting exposure assessments specifically related to 2,4-D. Moreover, the majority of the studies involved occupational exposures to a wide variety of chemical, physical, and biological agents including phenoxy herbicides, and it was difficult to discern specific exposure to 2,4-D. Without specific information regarding exposure to 2,4-D and with the contribution

Chapter | 84  Phenoxy Herbicides (2,4-D)

of other confounding factors, the establishment of a doseresponse relationship is difficult to ascertain. In cohort studies, exposure to 2,4-D could be reasonably assumed; however, no conclusive evidence was reported to show an association between 2,4-D and cancer. The only positive correlations were reported in three large cohort studies (Saracci et al., 1991; Wigle et al., 1990; Wiklund and Holm, 1986; Wiklund et al., 1987), in which exposures were primarily to 2,4,5-T or were mixed with other herbicides, and none of the reported effects were consistent among the studies. Although there has been a persistent hypothesis that exposure to 2,4-D may be associated with an increased incidence of non-Hodgkin’s lymphoma (Hoar et al., 1986; Zahm et al., 1990), the analysis by Munro et al. (1992) concluded that the results of these case-control studies did not strongly support the hypothesis based on weaknesses in the methodology employed and lack of control for other possible risk factors for non-Hodgkin’s lymphoma (e.g., viruses and immune system modulation). Other reviews have been conducted since the work by Munro et al. (1992). A SAB/SAP Special Joint Committee from the U.S. Environmental Protection Agency (EPA, 1994) reviewed the available data on 2,4-D and concluded that “… the data are not sufficient to conclude that there is a cause and effect relationship between the exposure to 2,4-D and non-Hodgkin’s lymphoma” and 2,4-D still remains classified as a Group D (not classifiable as to human carcinogenicity) (EPA, 1997). Similarly, the Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment and the WHO Expert Group on Pesticide Residues reviewed the data on 2,4-D and stated that the results of the available epidemiology studies are inconsistent and that any reported associations are weak (Rowland, 1997). The National Cancer Institute of Canada also convened an Ad Hoc Panel on Pesticides and Cancer which concluded that “it was not aware of any definitive evidence to suggest that synthetic pesticides contribute significantly to overall cancer mortality” (Ritter, 1997). A few additional studies have been published since the above reviews (Becher et al., 1996; Burns et al., 2001; Fleming et al., 1997; Zahm, 1997). Three of these are mortality studies in factory workers (Becher et al., 1996; Burns et al., 2001) and lawn care workers (Zahm, 1997), and one is a retrospective cohort study in pesticide applicators (Fleming et al., 1997). As with previous epidemiology studies, the findings were inconsistent and did not provide any conclusive evidence of increased cancer risk associated with exposure to 2,4-D. In fact, Fleming et al. (1997) reported that in a cohort of 33, 669 pesticide applicators, there were no confirmed cases of soft tissue sarcoma or non-Hodgkins lymphoma and Burns et al. (2001) concluded there was “no evidence of causal association between exposure to 2,4-D and mortality due to all causes and malignant neoplasm” and “no significant risk due to NHL was found.” In addition, Acquavella et al. (1998)

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conducted a meta-analysis of 37 studies in farmers and concluded that farmers did not have elevated rates of several cancers, with the exception of lip cancer. In a preliminary study with a small group of farmers, Faustini et al. (1996) reported that 2,4-D exposure affected some immunological variables; however, the data were highly variable (i.e., large standard deviations) and the group tested was very small (n  10). There has been some discussion in the literature linking immunosuppressive agents with an increased incidence of non-Hodgkin’s lymphoma (Hardell et al., 1998; Hardell and Axelson, 1998), but there has been no clear or consistent evidence in humans indicating that 2,4-D affects the immune system. This is well supported by animal studies.

84.8  Summary 2,4-D is the most common of the phenoxy herbicides and is one of the best-studied agricultural chemicals. It is primarily used as a herbicide in agriculture, forestry, and lawn care practices, and is effective against a wide variety of broadleaf plants. Occupational exposure to 2,4-D is mainly through dermal contact but can also occur, to a lesser extent, via ingestion and inhalation. Studies in humans have shown exposures to be extremely low even in all occupational groups (i.e., 40 g/kg body weight/day). The extensive database of metabolic, toxicological, and epidemiological studies on 2,4-D has provided no evidence that 2,4-D poses any health risk to humans when used according to label directions.

References Acquavella, J., Olsen, G., Cole, P., Ireland, B., Kaneene, J., Schuman, S., and Holden, L. (1998). Cancer among farmers: A meta-analysis. Ann. Epidemiol. 8(1), 64–74. Agriculture Canada (1983). “Re: 2,4-D Products Registered Under the Authority of the Pest Control Products Act,” (Memorandum to Registrants) No. R-1-216. Agriculture Canada, Food Production and Inspection Branch, Ottawa, Ontario. Alexander, B. H., Marndel, J. S., Baker, B. A., Burns, C. J., Bartels, M. J., Acquavella, J. F., and Gustin, C. (2007). Biomonitoring of 2,4-dichlorophenoxyacetic acid exposure and dose in Farm Families. Env. Hlth. Perspect 115, 370–376. Anderson, D., and Styles, J. A. (1978). The bacterial mutation test. Br. J. Cancer 37, 924–930. Arkhipov, G. N., and Kozlova, I. N. (1974). A study of the carcinogenic potential of a herbicide: 2,4-D amine salt. Voprosy Pitaniia 5, 83–84. Arnold, E. K., Beasley, V. R., Parker, A. J., and Stedelin, J. R. (1991). 2,4-D toxicosis. II. A pilot study of clinical pathologic and electroencephalographic effects and residues of 2,4-D in orally dosed dogs. Vet. Hum. Toxicol. 33(5), 446–449. Aylward, L. L., Morgan, M. K., Arbuckle, T. E., Barr, D. B., Burns, C. J., Alexander, B. H., and Hays, S. M. (2009). Biomonitoring data for 2,4-dichlorophenoxyacetic acid in the US and Canada: Interpretation

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in a public health context using Biomonitoring Equivalents. Env. Hlth. Perspect. Online 12 Aug 2009, doi: 10.1289/ehp.0900970. Beasley, V. R., Arnold, E. K., Lovell, R. A., and Parker, A. J. (1991). 2,4-D toxicosis. I. A pilot study of 2,4-dichlorophenoxyacetic acidand dicambainduced myotonia in experimental dogs. Vet. Hum. Toxicol. 33(5), 435–440. Becher, H., Flesch-Janys, D., Kauppinen, T., Kogevinas, M., Steindorf, K., Manz, A., and Wahrendorf, J. (1996). Cancer mortality in German male workers exposed to phenoxy herbicides and dioxins. Cancer Causes Control 7, 312–321. Bernard, P. A., Toyoshima, E., Eccles, C. U., Mayer, R. F., Johnson, K. P., and Max, S. R. (1985). 2,4-Dichlorophenoxyacetic acid (2,4-D) reduces acetyl-cholinesterase activity in rat muscle. Exper. Neurol. 87, 544–556. Berry, D. L. (1989). “Final Report of the Determination of Halogenated Dibenzo-p-dioxins and Dibenzofurans in 2,4-Dichlorophenoxyacetic Acid,” No. AL 89-030290. Analytical Sciences Laboratories, The Dow Chemical Company, Midland, MI. Bionetics Research Laboratory (1968). Evaluation of Carcinogenic, Teratogenic, and Mutagenic Activities of Selected Pesticides and Industrial Chemicals. Carcinogenic Study; Volume II. Terato­genic Study in Mice and Rats. National Cancer Institute (NCI), Bethesda, MD. Blakley, B. R. (1986). The effect of oral exposure to the N-butylester of 2,4-dichlorophenoxyacetic acid on the immune response in mice. Int. J. Immunopharmacol. 8(1), 93–99. Blakley, B. R., and Blakley, P. M. (1986). The effect of prenatal exposure to the n-butylester of 2,4-dichlorophenoxyacetic acid (2,4-D) on the immune response in mice. Teratology 33, 15–20. Blakley, B. R., and Schiefer, B. H. (1986). The effect of topically applied n-butylester of 2,4-dichlorophenoxyacetic acid on the immune response in mice. J. Appl. Toxicol. 6(4), 291–295. Breslin, W. J., Liberacki, A. B., and Yano, B. L.(1991). “Isopropylamine Salt of 2,4-D: Oral Gavage Teratology Study in New Zealand White Rabbits,” Unpublished Report No. M-004725-013. Dow Chemical Company, Midland, MI. Burns, C. J., Beard, K. K., and Cartmill, J. B. (2001). Mortality in chemical workers potentially exposed to 2,4-dichlorophenoxyacetic acid (2, 4-D) 1945-94: an update. Occup. Environ. Med. 58(1), 24–30. Buslovich, S. Y. and Pichugin, Y. I. (1983). Electromyographic characteristics of acute poisonings with chlorine derivatives of phenoxy acids. Farmakologiia i Tosikologiia. 46(3), 99–101. Canadian Centre for Toxicology (CCT)(1987). “Panel Report on Carcinoge­ nicity of 2,4-D,” Canadian Centre for Toxicology (CCT), Guelph, ON. Carreon, R. et al., (1983). “2,4-Dichlorophenoxyacetic Acid Isopropyl­ amine Salt: Acute Toxicological Properties. Industry Task Force II on 2,4-D Research Data,” Unpublished Report. Dow Chemical Company, Midland, MI. Carreon, R., and Rao, K. S. (1985). “DMA-6 Weed Killer: Dermal Sensitization Potential in the Guinea Pig. Industry Task Force II on 2,4-D Research Data,” Unpublished Report. Dow Chemical Company, Midland, MI. Charles, J. M., Bond, D. M., Jeffries, T. K., Yano, B. L., Stott, W. T., Johnson, K. A., Cunny, H. C., Wilson, R. D., and Bus, J. S. (1996a). Chronic dietary toxicity/oncogenicity studies on 2,4-dichlorophenoxyacetic acid in rodents. Fund. Appl. Toxicol. 33, 166–172. Charles, J. M., Cunny, H. C., Wilson, R. D., and Bus, J. S. (1996b). Comparative subchronic studies on 2,4-dichlorophenoxyacetic acid, amine and ester in rats. Fund. Appl. Toxicol. 33, 161–165. Charles, J. M., Dalgard, D. M., Cunny, H. C., Wilson, R. D., and Bus, J. S. (1996c). Comparative subchronic and chronic dietary toxicity

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studies on 2,4-dichlorophenoxyacetic acid, amine and ester in the dog. Fund. Appl. Toxicol. 29, 78–85. Chernoff, N., Woodrow Setzer, R., Miller, D. B., Rosen, M. B., and Rogers, J. M. (1990). Effects of chemically induced maternal toxicity on prenatal development in the rat. Teratology 42, 651–658. Clark, D. E., Palmer, J. S., Radeleff, R. D., Crookshank, H. R., and Farr, F. M. (1975). Residues of chlorophenoxy acid herbicides and their phenolic metabolites in tissues of sheep and cattle. J. Agric. Food Chem. 23(3), 573–578. Clausen, M., Leier, G., and Witte, I. (1990). Comparison of the cytotoxicity and DNA-damaging properties of 2,4-D and U 46 D fluid (dimethylammonium salt of 2,4-D). Arch. Toxicol. 64, 497–501. Cochrane, W. P., Singh, J., Miles, W., and Wakeford, B. (1981). Determination of chlorinated dibenzo-p-dioxin contaminants in 2,4-D products by gas chromatography-mass spectrometric techniques. J. Chromatogr. 217, 289–299. Cochrane, W. P., Singh, J., Miles, W., and Wakeford, B. (1982a). “Levels of Polychlorinated Dibenzo-p-Dioxins in 2,4-D Technical Products,” Agriculture Canada, Food Production and Inspection Branch, Laboratory Services Division, Ottawa (ON). Cochrane, W. P., Singh, J., Miles, W., Wakeford, B., and Scott. (1982b). Analysis of technical and formulated products of 2,4-dichlorophenoxyacetic acid for the presence of chlorinated dibenzo- p-dioxins. In: Impact of Chlorinated Dioxins and Related Compounds on the Environment. Oct. 22–24, 1980, Rome. pp. 1–5. Collins, T. F. X., and Williams, C. H. (1971). Teratogenic studies with 2,4,5-T and 2,4-D in the hamster. Bulletin Environ. Contamin. Toxicol. 6(6), 559–567. Courtney, K. D. (1977). Prenatal effects of herbicides: Evaluation by the prenatal development index. Arch. Environ. Contam. Toxicol. 6, 33–46. Cramer, P. H. (1996). “2,4-Dichlorophenoxyacetic Acid: Analysis for Polychlorinated Dibenzo-p-dioxins and Dibenzofurans in Acid, Salt, and/or Ester Technical Material,” Lab Project ID MRI No. 4199. Midwest Research Institute, Kansas City, MO, for DowElanco, Indianapolis, IN. Crossen, P. E., Morgan, W. F., Horan, J. J., and Stewart, J. (1978). Cytogenetic studies of pesticide and herbicide sprayers. N. Zeal. Med. J. 88(619), 192–195. de Duffard, A. M. E., de Alderete, M. N., and Duffard, R. (1990a). Changes in brain serotonin and 5-hydroxyindolacetic acid levels induced by 2,4-dichlorophenoxyacetic butyl ester. Toxicology 64, 265–270. de Duffard, A. M. E., Orta, C., and Duffard, R. (1990b). Behavioral changes in rats fed a diet containing 2,4-dichlorophenoxyacetic butyl ester. Neurotoxicology 11(4), 563–572. Duffard, R., Bortolozzi, A., Ferri, A., Garcia, G., and Evangelista de Duffard, A. M. (1995). Developmental neurotoxicity of the herbicide 2,4-dichlorophenoxyacetic acid. Neurotoxicology 16(4), 764. Duggan, R. E., and Corneliussen, P. E. (1972). Dietary intake of pesticide chemicals in the United States (III), June 1968–April 1970. Pestic, monit. J., 5(4) 331–341. (Cited in WHO, 1984). Duggan, R. E., and Lipscomb, G. Q. (1969). Dietary intake of pesticide chemicals in the United States (II), June 1966-April 1968. Pestic. Monit. J. 2(4), 153–162. Elo, H., Hervonen, H., and Ylitalo, P. (1988). Comparative study on cerebrovascular injuries by three chlorophenoxyacetic acids (2,4-D, 2,4,5T and MCPA). Comp. Biochem. Physiol. 90C(1), 65–68. Elo, H. A., and Macdonald, E. (1989). Effects of 2,4-dichlorophenoxyacetic acid (2,4-D) on biogenic amines and their acidic metabolites in brain and cerebrospinal fluid of rats. Arch. Toxicol. 63, 127–130.

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in Food—1996: Toxicological Evaluations.” WHO/PCS/97 1, 45–96. Food and Agriculture Organization of the United Nations (FAO), Rome, Italy. Rudel, R., and Senges, J. (1972). Experimental myotonia in mammalian skeletal muscle: Changes in membrane properties. Pflugers. Arch. 331, 324–334 (Cited in Bernard et al., 1985). Saracci, R., Kogevinas, M., Bertazzi, P. A., Demesquita, B. H. B., Coggon, D., Green, L. M., Kauppinen, T., Labbe, K. A., Littorin, M., Lynge, E., Mathews, J. D., Neuberger, M., Osman, J., Pearce, N., and Winkelmann, R. (1991). Cancer mortality in workers exposed to chlorophenoxy herbicides and chlorophenols. Lancet 338(8774), 1027–1032. Sauerhoff, M. W., Braun, W. H., Blau, G. E., and Gehring, P. J. (1977). The fate of 2,4-dichlorophenoxyacetic acid (2,4-D) following oral administration to man. Toxicology 8, 3–11. Schultz, S. K., Brock, A. W., and Killeen, J. C.(1990). “Dermal Sensitization Study (Closed Patch Repeated Insult) in Guinea Pigs, Rabbits with Diethanolamine Salt of 2,4-D,” Unpublished Report 90–165. Industry Task Force II on 2,4-D Research Data. Ricerca Inc, Painesville, OH. Schulze, G. E. (1990a). “21-Day Dermal Irritation and Dermal Toxicity Study in Rabbits with 2,4-Dichlorophenoxyacetic Acid,” HLA Study No. 2184-109. Hazleton Laboratories America Inc, Vienna, VA. Schulze, G. E. (1990b). “21-Day Dermal Irritation and Dermal Toxicity Study in Rabbits with Dimethylamine Salt of 2,4Dichlorophenoxyacetic Acid,” HLA Study No. 2184-111. Hazleton Laboratories America Inc, Vienna, VA. Schulze, G. E. (1990c). “21-Day Dermal Irritation and Dermal Toxicity Study in Rabbits with 2,4-Dichlorophenoxyacetic Acid-2-Ethylhexyl Ester,” HLA Study No. 2184-110. Hazleton Laboratories America Inc, Vienna, VA. Schulze, G. E. (1991). “Final Report: Subchronic Toxicity Study in Mice with 2,4-Dichlorophenoxyacetic Acid,” Industry Task Force II on 2,4-D Research Data. Hazleton Laboratories America, Inc, Vienna, VA. Schulze, G. E., and Dougherty, J. A. (1988). Neurobehavioral toxicity of 2,4-D-n-butyl ester (2,4-D ester): Tolerance and lack of cross-tolerance. Neurotoxicol. Teratol. 10, 75–79. Schwetz, B. A., Sparschu, G. L., and Gehring, P. J. (1971). The effect of 2,4-dichlorophenoxyacetic acid (2,4-D) and esters of 2,4-D on rat embryonal, foetal and neonatal growth and development. Fd. Cosmet. Toxicol. 9, 801–817. Sell, C. R., Maitlen, J. C., and Aller, W. A. (1982). Perspiration as an important physiological pathway for the elimination of 2,4-dichlorophenoxyacetic acid from the human body. Am. Chem. Soc. Abstr. Pap. 183 PEST #74. Serota, D. G. (1983a). “Subchronic Toxicity Study in Rats with 2,4-D Acid,” Unpublished Report 2184-102. Industry Task Force II on 2,4-D Research Data. Hazleton Laboratories America, Inc, Vienna, VA. Serota, D. G. (1983b). “Subchronic Toxicity Study in Mice with 2,4-D Acid,” Unpublished Report 2184-100. Industry Task Force II on 2,4-D Research Data. Hazleton Laboratories America, Inc, Vienna, VA. Serota, D. G. (1986). “Combined chronic toxicity and oncogenicity study in rats with 2,4-D acid,” Unpublished Report 2184-103. Industry task Force II on 2,4-D Research Data. Hazleton Laboratories America, Inc, Vienna, VA. Serota, D. G. et al. (1986). “Combined Chronic Toxicity and Oncogenicity Study in Rats with 2,4-D Acid,” Unpublished Report, 2184-102. Hazleton Laboratories America, Inc, Vienna, VA. Serota, D. G. et al. (1987). “Oncogenicity Study in Mice with 2,4-D Acid,” Unpublished Report, 2184-101. Hazleton Laboratories America, Inc, Vienna, VA.

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Serrone, D. M., Killeen, J. C., and Benz, G. (1991). “A Subchronic Toxicity Study in Rats with the Diethanolamine Salt of 2,4Dichlorophenoxyacetic Acid,” Unpublished Report 90-0186. Ricera Inc, Painesville, OH. Simmon, V. F., Kauhanen, K., and Tardiff, R. G. (1977). Mutagenic activity of chemicals identified in drinking water. In Progress in Genetic Toxicology (B. A. Bridges Scott and F. H. Sodels, eds.) Vol. 2, pp. 249–258. Elsevier/North-Holland, Amsterdam. Soler-Niedziela, L., Ong, T., Nath, J., and Zeiger, E. (1988). Mutagenicity studies of dioxin and related compounds with the Salmonella arabinose resistant assay system. Toxic. Assess 3(2), 137–145. Steiss, J. E., Braund, K. G., and Clark, E. G. (1987). Neuromuscular effects of acute 2,4-dichlorophenoxyacetic acid (2,4-D) exposure in dogs. J. Neurol. Sci. 78, 295–301. Styles, J. A. (1973). Cytotoxic effects of various pesticides in vivo and in vitro. Mutat. Res. 21(1), 50–51. Styles, J. A. (1977). Mammalian cell transformation in vitro. Br. J. Cancer 37, 931–936. Szabo, J. R., and Rachunek, B. L. (1991). “2,4-D, Butoxyethyl Ester: 13Week Dietary Toxicity Study in Fischer Rats,” ID. DECO-TXT:K007722-015. Dow Elanco, Indianapolis, IN. Thomas, K. W., Dosemeci, M., Hoppin, J. A., Sheldon, L. S., Croghan, C. W., Gordon, S. M., Jones, M. A., Reynolds, S. J., Raymer, J. H., Akland, G. C., Lynch, C. F., Knott, C. E., Sandler, D. P., Blair, A. E., and Alavanja, M. C. (25 Feb, 2009). Urinary biomarker, dermal, and air measurement results for 2,4-D and chlorpyrifos farm applicators in the Agricultural Health Study. J. Exp. Sci. Env. Epidemiol. Advance online publication; doi:10.1038/jes.2009.6. Toyoshima, E., Mayre, R. F., Max, S. R., and Eccles, C. (1985). 2,4Dichlorophenoxyacetic acid (2,4-D) does not cause polyneuropathy in rats. J. Neurol. Sci. 70, 225–229. Tyynela, K., Elo, H. A., and Ylitalo, P. (1990). Distribution of three common chlorophenoxyacetic acid herbicides into the rat brain. Arch. Toxicol. 64(1), 61–65. Tyynela, K., Elo, H., Ylitalo, P., and Hervonen, H. (1986). The central nervous system toxicity of chlorophenoxyacetic acid herbicides. Arch. Toxicol. (Suppl. 9), 355. Unger, T. M., Kliethermes, J., Van Goethem, D., and Short, R. D.(1981). “Teratology and Postnatal Studies in Rats of the Propylene Glycol Butyl Ether and Isooctyl Esters of 2,4-Dichlorophenoxyacetic Acid,” PB8, 1, 191140. U.S. Environmental Protection Agency, Research Triangle Park, NC. Waters, G. D., Simmon, V. F., Mitchell, A. D., Jorgenson, T. A., and Valencia, R. (1980). An overview of short-term tests for the mutagenic and carcinogenic potential of pesticides. J. Environ. Sci. Health B15(6), 867–906. Wigle, D. T. et al. (1990). Mortality study of Canadian male farm operators: Non-Hodgkin’s lymphoma mortality and agricultural practices in Saskatchewan. J. Nat. Cancer Inst. 82(7), 575–582. Wiklund, K., and Holm, L. E. (1986). Soft tissue sarcoma risk in Swedish agricultural and forestry workers. J. Nat. Cancer Inst. 76(2), 229–234. Wiklund, K. et al. (1987). Soft tissue sarcoma risk among agricultural and forestry workers in Sweden. Chemosphere 16(8/9), 2107–2110. Wolfe, W. H. (1983). The epidemiology and toxicology of Agent Orange. In “Proceedings of the 14th Conference on Environmental Toxicology,” Dayton, OH. Woolson, E. A., Thomas, R. F., and Ensor, P. D. J. (1972). Survey of polychlorodibenzo-p-dioxin content in selected pesticides. J. Agr. Food Chem. 20(2), 351–354.

Chapter | 84  Phenoxy Herbicides (2,4-D)

World Health Organization (WHO) (1975). “Evaluations of Some Pesticides in Food,” The Monographs, WHO Pesticide Series No. 4, pp. 159-183. World Health Organization, Geneva. World Health Organization (WHO) (1984). 2,4-Dichlorophenoxyacetic Acid (2,4-D). Environmental Health Criteria 29, pp. 1–151. IPCS International Programme on Chemical Safety, United Nations Environment Programme, International Labour Organisation, and the World Health Organisation. Yano, B. L., Cosse, P. F., Atkin, L., and Corley, R. A.(1991a). “2,4-D Isopropylamine Salt (2,4-D IPA): A 13-Week Dietary Toxicity Study in Fischer 344 Rats,” ID. HET m-004725-006. Dow Elanco, Indianapolis, IN. Yano, B. L., Cosse, P. F., Markham, D. A., and Atkin, L.(1991b). “2,4-D Triisopropanolamine Salt (2,4-D IPA): A 13-Week Dietary Toxicity Study in Fischer 344 Rats,” ID. K-008866-006. DowElanco, Indianapolis, IN. Yeary, R. A. (1986). Urinary excretion of 2,4-D in commercial lawn specialists. Appl. Ind. Hyg. 1(3), 119–121. Ylitalo, P., Narhi, U., and Elo, H. A. (1990). Increase in the acute toxicity and brain concentrations of chlorophenoxyacetic acids by probenicid in rats. Gen. Pharmacol. 21(5), 811–814.

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Yoder, J., Watson, M., and Benson, W. W. (1973). Lymphocyte chromosome analysis of agricultural workers during extensive occupational exposure to pesticides. Mutat. Res. 21, 335–340. Zablotny, C. L., Yano, B. L., and Breslin, W. J., (1991). “2,4-D Triiso­ propanolamine Salt (2,4-D TIPA): A 13-Week Dietary Toxicity Study in Fischer 344 Rats,” Unpublished Report No., K, 008866-006. Zahm, S. H. et al. (1990). A case-control study of non-Hodgkin’s lymphoma and the herbicide 2,4-dichlorophenoxyacetic acid (2,4-D) in eastern Nebraska. Epidemiology 1(5), 349–356. Zahm, S. H. (1997). Mortality study of pesticide applicators and other employees of a lawn care service company. J. Occup. Environ. Med. 39(11), 1055–1067. Zetterberg, G. (1978). Genetic effects of phenoxy acids on microorganisms. Ecol. Bull. 27, 193–204. Zetterberg, G., Busk, L., Elovson, R., Starec-Nordenhammar, I., and Ryttman, H. (1977). The influence of pH on the effects of 2,4-D (2,4dichlorophenoxyacetic acid, Na salt) on the Saccharomyces cerevisiae and Salmonella typhimurium. Mutat. Res. 42, 3–18. Zhamsaranova, S. D., Lebedeva, S. N., and Lyashenko, V. A. (1987). The immunodepressive effects of the herbicide 2,4-D in mice. Gig. I. Sanit. 5, 80–81.

Chapter 85

Dicamba Paul R. Harp Lewisville, North Carolina

Chemical name 3,6-Dichloro-2-methoxybenzoic acid or 3,6-dichloroo-anisic acid. Structure

O

CI

OH

O CH3

crystalline solid. The chemical formula for dicamba is C8H6Cl2O3 and the molecular weight is 221.04. Dicamba has a melting point of 114–116°C and a calculated vapor pressure of 1.67 mPa at 25°C. Dicamba is resistant to hydrolysis and oxidation under normal environmental conditions and may remain in soils for 7–10 months. It is soluble in water (solubility of 6.1 g/l at 25°C) and will readily leach into runoff water [National Research Council (NRC), 1977; Tomlin, 2003].

CI FIGURE 85.1

85.3  Formulations and uses

85.1  Synonyms

Dicamba is a benzoic acid herbicide that acts by mimicking the effects of auxins (i.e., natural plant growth hormones), causing enhanced but uncontrolled growth rates, alterations in plant function homeostasis, and death. It was first registered for use in the United States in 1967. Dicamba is used to control a wide spectrum of annual and perennial broadleaf weeds and is effective in both pre- and postemergence applications. Primary agricultural applications include weed control in grain/cereal crops. Industrial/commercial applications include maintenance of pastures, forest lands, fence rows, and transportation and utility rights-of-way. Residential applications include weed control for lawns and golf courses. Commercial applications are formulated as liquids, granules, or water dispersible granules, whereas residential applications are marketed as granulated “weed-and-feed” products, liquid concentrates, or ready-to-use sprays. At the time of the most recent re-registration, the U.S. Environmental Protection Agency (EPA) reported more than 400 active products containing dicamba (U.S. EPA, 2006).

The common name is dicamba (ANSI, BSI, ISO). Code numbers include CAS 1918-00-9, CAS 1982-69-0 (dicamba sodium salt), CAS 2300-66-5 (dicamba dimethylamine salt), CAS 104040-79-1 (dicamba diglycolamine salt), CAS 55871-02-8 (dicamba isopropylamine salt), CAS 10007-85-9 (dicamba potassium salt), EINECS 217-635-6, EPA Pesticide Chemical Code 029801, and Velsicol 58-CS-11. Selected trade names of formulations containing dicamba as the sole active ingredient include Banvel, Diablo, Rifle, Sterling Blue, and Vanquish. Dicamba is also marketed in combination with other herbicides, such as 2,4-D, atrazine, carfentrazone-ethyl, clopyralid, dichlorprop, diflufenzopyr, fluroxypyr, glyphosate, MCPA, mecopropp, picloram, quinclorac, sulfentrazone, and others. Examples of dicamba premix formulations with other herbicides include Banlene Plus, Banlene Solo, Brushmaster, Diptyl, Distinct, Docklene, Durtok, GlyKamba, Marksman, Northstar, Novertex gazons H, Q4, Resolve, Selectone G, Surge, Trimec 992, Trimec Bentgrass, Trimec Classic, Trimec Encore, Trimec Plus, Trimec Southern, Trimonal, Tri-Power, Veteran 720, and Weedmaster.

85.2  Physical and chemical properties Dicamba in pure form is a colorless crystalline solid; technical-grade dicamba (~85% purity) is a buff-colored Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

85.4  Toxicokinetics Dicamba is readily absorbed following ingestion but only minimal absorption occurs after dermal exposure. Most animal studies indicate that greater than 90% of the original dose undergoes urinary excretion, with small amounts 1849

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excreted in the feces. Dicamba is excreted primarily in the free unmetabolized (parent) form but may also be conjugated with glucuronic acid. Elimination occurs rapidly, and there is no evidence of bioaccumulation in mammalian systems.

85.5  Toxicity to laboratory animals Dicamba has a low mammalian toxicity. Reported acute oral LD50 values are 0.757 to greater than 2.740 g/kg in rats, 0.566–3.000 g/kg in guinea pigs, and 0.566–2.000 g/kg in rabbits (Hayes, 1982; U.S. EPA, 2005). In rats, acute dermal LD50 values are greater than 2000 mg/kg and acute inhalation LC50 values are greater than 5.3 mg/l (U.S. EPA, 2005). Dicamba was reported to have weak promoting activity when coadministered in the diet with phenobarbital in a two-stage hepatocarcinogenesis study in female rats (Espandiari et al., 1999). The rats received a single initial dose of diethylnitrosoamine followed by dietary administration of dicamba, phenobarbital, or a combination for 6 months. No evidence of promotion was reported in the animals that received only dicamba in their diet. Dicamba, either alone or combined with pendimethalin (a preemergent herbicide) or 2,4-D and atrazine (postemergent herbicides), was shown to induce significant levels of apoptosis in a mouse preimplantation embryo assay, but none of the treatment combinations significantly affected either the mean cell number per embryo or the percentage of embryos developing to the blastocyst stage (Greenlee et al., 2004). In the same investigation, dicamba in combination with 2,4-D and MCPP (a mixture representing a commercial herbicide formulation used for lawn care) significantly reduced both the mean cell number per embryo and the percentage of embryos developing to the blastocyst stage, in the absence of a significant apoptotic effect. The U.S. EPA issued a re-registration eligibility decision for dicamba in 2006 (U.S. EPA, 2006). In brief, dicamba was considered to have low acute toxicity by oral, dermal, and inhalation routes. It was considered to be both an eye and a skin irritant but not a dermal sensitizer. The U.S. EPA reported that signs of neurotoxicity were noted in numerous studies conducted in rats and rabbits in groups receiving high doses of dicamba. Exposures to dicamba in utero and/or pre-/postnatal in rats and rabbits did not indicate any potential for the herbicide to act as a developmental toxicant. The U.S. EPA acknowledged reports of positive findings in some studies but considered dicamba to be nonmutagenic based on negative findings in numerous endpoints, including mutations (in Ames Salmonella assay), chromosome aberrations in Chinese hamster ovary (CHO) cells, and unscheduled DNA synthesis (in human lung fibroblasts). The U.S. EPA classified dicamba as “not likely to be carcinogenic to humans” by the oral route based on lack of evidence of effects in rat and mouse carcinogenicity studies (U.S. EPA, 2005).

Hayes’ Handbook of Pesticide Toxicology

85.6  Toxicity to humans Evaluation of the specific adverse effects of dicamba in humans has been confounded by the relatively small amount of data available from exposures exclusively to dicamba. Many of the documented exposures have been to products containing a combination of herbicidal agents. Signs of dicamba intoxication include vomiting, breathing difficulty, cyanosis, depression, loss of appetite, weight loss, incontinence, and muscular weakness/exhaustion subsequent to muscle spasms (myotonia). Dicamba is a skin/eye irritant and may also be mildly corrosive. Effects of dermal/ocular exposure are generally temporary, but severe or permanent ocular damage may be possible and appropriate eye protection should be used, especially when working with concentrated solutions.

85.7  Reproductive effects No adverse reproductive effects were reported in threegeneration reproductive studies conducted in rats fed diets containing dicamba at levels up to 500 ppm (NRC, 1977).

85.8  Genotoxic effects During the 1970s and 1980s, numerous researchers employed various assays in investigation of the mutagenic potential of dicamba, with somewhat contradictory and inconclusive results (references cited in Perocco et al., 1990). More recent efforts to better characterize the potential genotoxicity of dicamba are summarized here. Perocco and co-workers (1990) reported a DNA-damaging potential of dicamba based on an enhanced unwinding rate of rat liver DNA following in vivo exposure to dicamba (single intraperitoneal injection). In in vitro experiments with human peripheral blood lymphocytes, Perocco and coworkers also found dicamba to induce unscheduled DNA synthesis (in the presence but not in the absence of S9 metabolic activation) and slightly increase the frequency of sister chromatid exchanges (SCE; in both the presence and the absence of S9 metabolic activation). In no case did the dicamba-induced increase in SCE exceed two times the spontaneous frequency of exchanges, so the positive response in this endpoint, although statistically significant, was not considered to be definitive. In a follow-up study by some of the same researchers, dicamba was also found to be nonclastogenic in an in vivo bone marrow chromosome aberration study (Hrelia et al., 1994). Dicamba has been shown to induce peroxisomal enzymes in rat liver and cause transcriptional upregulation of the peroxisome proliferator-activated receptor (Espandiari et al., 1995). Although tumorigenic effects of dicamba have not been reported, long-term exposure to peroxisome proliferators has been associated with increased hepatic tumor frequency in rodents. Further investigation also revealed activation of

Chapter | 85  Dicamba

hepatic nuclear factor-κB, a transcription factor that may be involved in the hepatocarcinogenic action of certain peroxisome proliferators (Espandiari et al., 1998). Dicamba has been reported to uncouple oxidative phosphorylation through a combination of inhibition of redox complexes and reduction in the integrity of the mitochondrial membrane, resulting in enhanced proton leakage (Peixoto et al., 2003). Using a novel transgenic plant-based system (Arabidopsis thaliana), Filkowski and co-workers reported dicamba-induced increases in the frequencies of point mutations and homologous recombination at microgram per liter concentrations. These events were observed at test article concentrations that did not produce visible physiological effects on the plants and were lower than those previously reported as cyto- or genotoxic in bacterial or mammalian cell culture systems (Filkowski et al., 2003). A series of studies were conducted by Gonzalez and coworkers in an attempt to better elucidate the potential for dicamba to cause genotoxicity and the mechanism(s) through which this might occur. These studies included both dicamba and a commercial herbicide formulation, Banvel (~58% dicamba). In the first study, conducted in in vitro wholeblood human lymphocyte cultures, both dicamba and Banvel were found to increase the frequency of sister chromatid exchanges (SCEs) and also result in cytotoxicity, as indicated by either delay in cell cycle progression or reduction in the proliferative rate index (Gonzalez et al., 2006). Similar types of genotoxic effects of dicamba and Banvel were subsequently demonstrated in CHO cells as measured by both SCE and single cell gel electrophoresis (SCGE) assays (Gonzalez et al., 2007). Frequency of SCE was increased in a non-dose-dependent manner across all tested concentrations of dicamba and Banvel (1–500 g/ ml for 24 h), with the maximum frequency of exchanges occurring with 200 and 500 g/ml dicamba and Banvel, respectively. Similarly, frequencies of slightly damaged and damaged cells, as measured by SCGE, increased in a non-dose-dependent manner following a 90-min exposure to dicamba or Banvel at concentrations ranging from 50 to 500 g/ml. In the same study, dicamba was found to induce cytotoxicity in CHO cells at 200 and 500 g/ml as measured by changes in cell cycle kinetics and at 500 g/ml as indicated by a reduction in the proliferative rate index (PRI). Banvel also altered cell cycle kinetics at 500 g/ml but did not cause a statistically significant reduction in PRI at any concentration tested. Further investigation by the same researchers found that both the increase in frequency of SCEs and the alterations in cell cycle kinetics induced in CHO cells by dicamba (500 g/ml) could be attenuated by -tocopherol (vitamin E, 25 M), suggesting a role of oxidative damage in the genotoxicity activity of this herbicide (Gonzalez et al., 2009). Collectively, the available information suggests that dicamba has both genotoxic and cytotoxic potential.

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Evidence of genotoxicity has been demonstrated by multiple researchers in in vitro systems in the absence of S9 metabolic fractions, suggesting that the effects are related to dicamba rather than a metabolite. Furthermore, the report of -tocopherol attenuation of dicamba-induced SCE suggests that the mechanism(s) of genotoxicity may include oxidative damage (Gonzalez et al., 2009).

85.9  Treatment of poisoning Treatment is symptomatic and supportive. No specific antidote is available.

Conclusion Dicamba is a widely used benzoic acid herbicide that acts by mimicking the effects of natural plant growth hormones. It is used to control a wide spectrum of broadleaf weeds and was first registered for use in the United States in 1967. Dicamba is readily absorbed after oral exposure and rapidly excreted, primarily in the urine in unmetabolized form. It has a low acute mammalian toxicity and is a skin/eye irritant but does not cause dermal sensitization. Some studies have indicated a potential genotoxic effect of dicamba, but chronic studies in rats and mice have not demonstrated evidence of carcinogenicity.

References Espandiari, P., Thomas, V. A., Glauert, H. P., O’Brien, M., Noonan, D., and Robertson, L. W. (1995). The herbicide dicamba (2-methoxy-3, 6-dichlorobenzoic acid) is a peroxisome proliferator in rats. Fundam. Appl. Toxicol. 26, 85–90. Espandiari, P., Ludewig, G., Glauert, H. P., and Robertson, L. W. (1998). Activation of hepatic NF-κB by the herbicide dicamba (2-methoxy3,6-dichlorobenzoic acid) in female and male rats. J. Biochem. Mol. Toxicol. 12, 339–344. Espandiari, P., Glauert, H. P., Lee, E. Y., and Robertson, L. W. (1999). Promoting activity of the herbicide dicamba (2-methoxy-3,6dichlorobenzoic acid) in two stage hepatocarcinogenesis. Int. J. Oncol. 14(1), 79–84. Filkowski, J., Besplug, J., Burke, P., Kovalchuk, I., and Kovalchuk, O. (2003). Genotoxicity of 2,4-D and dicamba revealed by transgenic Arabidopsis thaliana plants harbouring recombination and point mutation markers. Mutat. Res. 542, 23–32. Gonzalez, N. V., Soloneski, S., and Larramendy, M. L. (2006). Genotoxicity analysis of the phenoxy herbicide dicamba in mammalian cells in vitro. Toxicol. in Vitro 20, 1481–1487. Gonzalez, N. V., Soloneski, S., and Larramendy, M. L. (2007). The chlorophenoxy herbicide dicamba and its commercial formulation Banvel induce genotoxicity and cytotoxicity in Chinese hamster ovary (CHO) cells. Mutat. Res. 634, 60–68. Gonzalez, N. V., Soloneski, S., and Larramendy, M. L. (2009). Dicambainduced genotoxicity in Chinese hamster ovary (CHO) cells is prevented by vitamin E. J. Hazard. Mater. 163, 337–343.

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Greenlee, A. R., Ellis, T. M., and Berg, R. L. (2004). Low-dose agrochemicals and lawn-care pesticides induce developmental toxicity in murine preimplantation embryos. Environ. Health Perspect. 112, 703–709. Hayes, W. J. (1982). Herbicides. In “Pesticides Studied in Man,” p. 536. Williams & Wilkins, Baltimore. Hrelia, P., Vigagni, F., Maffei, F., Morotti, M., Colacci, A., Perocco, P., Grilli, S., and Cantelli-Forti, G. (1994). Genetic safety evaluation of pesticides in different short-term tests. Mutat. Res. 321, 219–228. National Research Council (1977). Dicamba. In “Drinking Water and Health, Vol. 1,” pp. 521–525. National Academy of Sciences, Washington, DC. Peixoto, F., Vicente, J. A. F., and Madeira, V. M. C. (2003). The herbicide dicamba (2-methoxy-3,6-dichlorobenzoic acid) interacts with mitochondrial bioenergetic functions. Arch. Toxicol. 77, 403–409.

Hayes’ Handbook of Pesticide Toxicology

Perocco, P., Ancora, G., Rani, P., Valenti, A. M., Mazzullo, M., Colacci, A., and Grilli, S. (1990). Evaluation of genotoxic effects of the herbicide dicamba using in vivo and in vitro test systems. Environ. Mol. Mutagen. 15, 131–135. Tomlin, C. D. S. (2003). Dicamba. In “The Pesticide Manual” (C. D. S. Tomlin ed.), pp. 278–280. British Crop Protection Council, Hampshire, UK. U.S. Environmental Protection Agency (EPA) (2005). “Dicamba: HED Chapter of the Reregistration Eligibility Decision Document (RED) – I,” Office of Prevention, Pesticides and Toxic Substances, U.S. EPA, Washington, DC. U.S. Environmental Protection Agency (EPA) (2006). “Reregistration Eligibility Decision for Dicamba and Associated Salts,” Office of Pesticide Programs, U.S. EPA, Washington, DC.

Chapter 86

Imidazolinones* Frederick G. Hess1, Jane E. Harris2, Kimberly Pendino and Kathryn Ponnock3 1

BASF Corporation; 2Hoffman-La Roche, Inc.; 3Middlesex County Community College

86.1  Identity, properties, and uses 86.1.1  Chemical Names Imazapyr  (Arsenal® herbicide): 2-(4-isopropyl-4-methyl5-oxo-2-imidazolin-2-yl) nicotinic acid. Imazamethabenz-methyl  (Assert® herbicide): methyl 2(4-isopropyl-4-methyl-5-oxo-2-imidazolin-2-yl)-p-toluate mixed with methyl 6-(4-isopropyl-4-methyl-5-oxo-2-imidazolin-2-yl)-m-toluate (3:2). Imazapic  (Cadre® herbicide): 2-(4-isopropyl-4-methyl5-oxo-2-imidazolin-2-yl)-5-methymicotinicacid. Imazethapyr  (Pursuit® herbicide): 5-ethyl-2-(4-isopropyl-4-methyl-5-oxo-2-imidazolin-2-yl) nicotinic acid. Imazamox  (Raptor® herbicide): 2-(4-isopropyl-4-methyl5-oxo-2-imidazolin-2-yl)-5-(methoxymethyl) nicotinic acid. Imazaquin  (Scepter® herbicide): 2-(4-isopropyl-4-methyl5-oxo-2-imidazolin-2-yl)-3-quinoline carboxylic acid.

The very low vapor pressures of the imidazolinones indicate a low potential to volatilize. With the exception of imazamethabenz-methyl, which is an ester, the imidazolinones have pKa values (dissociation constants) for the carboxylic acid group that range from 3.0 to 3.5. Based on these dissociation constants, both the water solubility and n-octanol/ water partition coefficients are pH dependent. For example, if the localized pH is above approximately 3.5, the amount of the imidazolinone in ionized form increases, which consequently increases water solubility and decreases the n-octanol/water partition coefficients. In contrast, if the local pH is below approximately 3.0, the amount of the imidazolinone in ionized form decreases, which consequently decreases water solubility and increases the n-octanol/water partition coefficients. In addition, the imidazolinone herbicides demonstrate other dissociation constants at approximately 2 and 11, which, in turn, alter the specific water solubility and n-octanol/water partition coefficients.

86.1.2  Physical and Chemical Properties

86.1.3  Structure

The physical and chemical properties of the six imidazolinone herbicides (technical products) are presented in Table 86.1. The respective empirical formulas, molecular weights (range from 261–311), physical state (off-white to tan powders), melting points (range from 144–222°C), and vapor pressures (1  107 mm Hg) are very similar. One might expect that such close similarities in physical properties would be reflected in similarities of toxicological profiles, as discussed in Section 86.2.

The chemical structures of the respective imidazolinone herbicides bear a close resemblance to each other, as presented in Fig. 86.1. The active molecule of each compound has an identical imidazolinone ring structure with a carboxylic acid group, or specifically a carboxylic ester group for imazamethabenz-methyl, attached to respective backbone groups, such as a pyridine ring. Several of the compounds show very similar structure– activity relationships. For instance, imazapic, imazethapyr, imazamox, and imazaquin differ only by the respective substituent groups that are attached to the pyridine ring (Fig. 86.1). Imazapic has a methyl group, whereas imazethapyr has an ethyl group, imazamox has a methoxymethyl group, and imazaquin has a benzene ring fused to the pyridine ring. One might expect that such close resemblance in chemical structures would be reflected in similarities of toxicological profiles, as discussed in Section 86.2.

*The summaries and evaluations contained in this chapter are, in most cases, based on unpublished proprietary data. A registration authority should not grant a registration on the basis of this information unless it has first received authorization for such use from the owner of the data or has received the data on which these summaries are based, either from the owner of the data or from a second party that has obtained permission from the owner of the data for this purpose.

Hayes’ Handbook of Pesticide Toxicology Copyright © 2001 Elsevier Inc. All rights reserved

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Table 86.1  Comparison of the Physical and Chemical Properties of the Imidazolinone Herbicides Technical herbicide

Properties Imazapyr

Imazamethabenz- Imazapic methyl

Imazethapyr

Imazamox

Imazaquin

Empirical formula

C13H15N3O3

C16H20N2O3

C14H17N3O3

C15H19N3O3

C15H19N4O3

C17H17N3O3

Molecular weight

261

288

275

289

305

311

Physical state

White-to-tan powder

Off-white powder

Off-white-to-tan powder

Light-tan powder

Off-white powder Light-tan powder

Melting point

169–173°C

144–153°C

204–206°C

177–179°C

166–167°C

219–222°C

Vapor pressure 1  107 mmHg 1.13  108 mmHg 1  107 mmHg 1  107 mmHg 1  107 mmHg 2  108 mmHg at 60°C at 25°C at 60°C at 60°C at 60°C at 45°C

86.1.4  History and Uses

86.2  Toxicity to laboratory animals

The unique class of synthetic chemical compounds called the imidazolinone herbicides was discovered in the 1970s, with the first U.S. patent awarded in 1980 for imazamethabenz-methyl (Assert®). Other imidazolinones in the series, imazapyr (Arsenal®), imazapic (Cadre®), imazethapyr (Pursuit®), and imazaquin (Scepter®), received U.S. patents in 1989. More recently, imazamox (Raptor®) received a U.S. patent in 1994. The active ingredients of the imidazolinone herbicides are effective for selective postemergent weed control for the following crops: cereals, including wheat and barley (imazamethabenz-methyl); peanuts (imazapic); soybeans (imazamox, imazaquin, and imazethapyr); and other legumes, including peas, beans, and alfalfa (imazamox and imazethapyr). In addition, imazapyr is a broad-spectrum herbicide that is effective for noncrop uses for total vegetation control on industrial sites and railroad, highway, and utility rights-of-way and for forestry applications. Most recently, the imidazolinone herbicides have shown specificity for postemergent weed control on imidazolinone-tolerant corn called Imi-Corn® (imazethapyr and imazapyr), Smart™ canola (imazethapyr and imazamox), imidazolinone-tolerant rice (imazethapyr), imidazolinonetolerant sugar beet (imazamox), and imidazolinone-tolerant wheat (imazamox). After application, each imidazolinone herbicide is taken up by the foliage and/or roots of the susceptible weed, and subsequently, becomes translocated throughout the plant. Susceptible weeds stop growing and competing with the specific crop shortly after translocation; the weeds die within several weeks postapplication.

Following extensive testing in the required mammalian toxicity studies, the six imidazolinone herbicides demonstrate a low toxicological potential. The results of these studies with the respective technical products are presented in Table 86.2. The imidazolinone herbicides demonstrate this very low toxicity profile in mammals because their herbicidal activity relies on their mode of action, that is, the inhibition of the specific plant enzyme, acetohydroxyacid synthase (AHAS). AHAS is an important biosynthetic enzyme in the formation of three branched-chain aliphatic amino acids, namely, isoleucine, leucine, and valine, and is found in plant but not mammalian tissues. This inhibition of AHAS disrupts protein synthesis and subsequently interferes with DNA synthesis and cell growth, eventually leading to cell death in the specific weed, as described in Section 86.1.

86.2.1  Basic Findings 86.2.1.1  Acute Toxicity Studies The results of the acute toxicity tests indicate that the technical products of the imidazolinone herbicides are generally relatively nontoxic (U.S. EPA Toxicity Category IV) by the oral and inhalation routes of administration and only slightly toxic (Category III) by the dermal route, according to the respective LD50 and LC50 values presented in Table 86.2. Further, the technical products of this class of chemical compounds are either non-irritating or only slightly irritating in the rabbit primary skin irritation studies. The skin sensitization studies in guinea pigs, conducted according

Chapter | 86  Imidazolinones

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COOH

imazapyr (ARSENAL ® herbicide (BAS 693 H)

N

N HN

Figure 86.1  Chemical structures of the imidazolinone herbicides.

O

CH3

COOCH3

COOCH3

+

N

imazamethabenz-methy1 (ASSERT ® herbicide) (BAS 712 H)

N

CH3

HN

HN O

O

CH3

COOH

imazapic (CADRE ® herbicide) (BAS 715 H)

N

N HN

O

COOH

CH3

imazethapyr (PURSUIT ® herbicide) (BAS 685 H)

N

N HN

O

CH3

COOH

O

imazamox (RAPTOR ® herbicide) (BAS 720 H)

N

N HN

O

COOH

imazaquin (SCEPTER ® herbicide) (BAS 725 H)

N

N HN

O ® Registered

Trademark of BASF. Corporation

to the method of Buehler, demonstrate that the technical products of the six imidazolinone herbicides are nonsensitizers (Table 86.2). Results from the rabbit primary eye irritation studies with the technical products of the imidazolinone herbicides ranged from no irritation (imazaquin) to slightly irritating (imazamefhabenz-methyl and imazamox) to moderately irritating (imazapic and imazethapyr), showing complete recovery by day 7 postdosing. Consequently, these five imidazolinone herbicides are classified into U.S. EPA Toxicity Category III or IV. Finally, the rabbit primary eye

irritation study with imazapyr demonstrated irreversible irritation (Category I) based on two out of six animals with scattered opacities at day 21. Thus, the labels for imazapyr’s formulated products recommend protective eye-wear, which should mitigate any potential for eye irritation during mixing/loading and application.

86.2.1.2  Short-Term/Subchronic Toxicity Studies 21-Day Dermal Studies in the Rabbit  All six imidazolinone herbicides showed no dermal or systemic toxicity

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Table 86.2  Comparison of Mammalian Toxicity Data and Genotoxicity Data for the Imidazolinone Herbicides Technical herbicide

Study Imazamethabenzmethyl

Imazapic

Imazethapyr

Imazamox

Imazaquin

5000 mg/kg (relatively nontoxic) (Category IV)

5000 mg/kg (relatively nontoxic) (Category IV)

5000 mg/kg (relatively nontoxic) (Category IV)

5000 mg/kg (relatively nontoxic) (Category IV)

5000 mg/kg (relatively nontoxic) (Category IV)

5000 mg/kg (relatively nontoxic) (Category IV)

Acute dermal toxicity

2000 mg/kg

2000 mg/kg

2000 mg/kg

2000 mg/kg

4000 mg/kg

2000 mg/kg

(rabbit) LD50

(slightly toxic) (Category III)

(slightly toxic) (Category III)

(slightly toxic) (Category III)

(slightly toxic) (Category III)

(slightly toxic) (Category III)

(slightly toxic) (Category III)

1.3 mg/l (slightly toxic) (Category III)

1.4 mg/l (slightly toxic) (Category III)

4.8 mg/l (relatively nontoxic) (Category IV)

3.3 mg/l (relatively nontoxic) (Category IV)

6.3 mg/l (relatively nontoxic) (Category IV)

5.7 mg/l (relatively nontoxic) (Category IV)

Primary dermal irritation (rabbit)

Slight irritation (Category III)

Nonirritating (Category IV)

Slight irritation (Category IV)

Slight irritation (Category IV)

Slight irritation (Category IV)

Slight irritation (Category III)

Primary eye irritation (rabbit)

Irreversible irritation (Category I)

Slight irritation (Category III)

Moderate irritation (Category III)

Moderate irritation (Category III)

Slight irritation (Category III)

Nonirritating (Category IV)

Dermal sensitization (guinea pig) (Buehler method)

Nonsensitizer

Nonsensitizer

Nonsensitizer

Nonsensitizer

Nonsensitizer

Nonsensitizer

400 mg/kg b.w./day

200 mg/kg b.w./day (HDT) (HDT)

1000 mg/kg b.w./day

1000 mg/kg b.w./day (HDT) (HDT)

1000 mg/kg b.w./day

1000 mg/kg b.w./day

(HDT)

(HDT)

Acute toxicity Acute oral toxicity (rat) LD50

Acute inhalation toxicity (rat) LC50 (analytical)

Short-term/subchronic toxicity 21-day dermal

(Rabbit) (HDT) No observable effect level 90-day dietary 10,000 ppm (HCT) (dog) (250 mg/kg b.w./day) No observable effect level 90-day dietary 20,000 ppm (HCT) (rat) (1740 mg/kg b.w./day) No observable effect level (based on FC data)

1000 ppm (87.5 mg/kg b.w./day) (based on FC data)

(HDT) —

10,000 ppm (HCT) (250 mg/kg b.w./day)

40,000 ppm (HCT) — (1370 mg/kg b.w./day) (based on FC data)

20,000 ppm (HCT) (1625 mg/kg b.w./day) (based on FC data)

10,000 ppm (HCT) (820 mg/kg b.w./day) (based on FC data)

20,000 ppm (HCT) 10,000 ppm (HCT) (1660 mg/kg b.w./day) (830 mg/kg b.w./day) (based on FC data) (based on FC data)

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Imazapyr

10,000 ppm (HCT) (250 mg/kg b.w./day)

1000 ppm (25 mg/kg b.w./day)

No observable adverse effect level  5000 ppm (135 mg/kg b.w./day) (based on FC data)

1000 ppm (25 mg/kg b.w./day)

Systemic toxicity 10,000 ppm (HCT) No observable No observable effect level (1500 mg/kg b.w./day) Adverse effect level  525 ppm (79 mg/kg b.w./day)

7000 ppm (HCT) (1135 mg/kg/b.w./day)

5000 ppm (750 mg/kg b.w./day)

Oncogenicity 10,000 ppm (HCT) 2100 ppm (HCT) No observable effect level (1500 mg/kg b.w./day) (315 mg/kg b.w./day)

7000 ppm (HCT) (1135 mg/kg b.w./day) (based on FC data)

10,000 ppm (HCT) (1500 mg/kg b.w./day)

7000 ppm (HCT) 4000 ppm (HCT) (1200 mg/kg b.w./day) (600 mg/kg b.w./day) (based on FC data)

250 ppm (12.5 mg/kg b.w./day)

20,000 ppm (HCT) (1030 mg/kg b.w./day) (based on FC data)

10,000 ppm (HCT) (500 mg/kg b.w./day)

20,000 ppm (HCT) 10,000 pm (HCT) (1165 mg/kg b.w./day) (500 mg/kg b.w./day) (based on FC data)

4000 ppm (HCT) (200 mg/kg b.w./day)

20,000 ppm (HCT) (1030 mg/kg b.w./day) (based on FC data)

10,000 ppm (HCT) (500 mg/kg b.w./day)

20,000 ppm (HCT) 10,000 ppm (HCT) (1165 mg/kg b.w./day) (500 mg/kg b.w./day) (based on FC data)

400 mg/kg b.w./day (HDT) 400 mg/kg b.w./day (HDT)

500 mg/kg b.w./day

500 mg/kg b.w./day

300 mg/kg b.w./day

300 mg/kg b.w./day

250 mg/kg b.w./day

500 mg/kg b.w./day

700 mg/kg b.w./day (HDT)

1000 mg/kg b.w./day (HDT)

900 mg/kg b.w./day (HDT)

500 mg/kg b.w./day (HDT)

300 mg/kg b.w./day

No observable 1000 mg/kg b.w./day adverse effect (HDT) level  1000 mg/kg b.w./ day (HDT) 1000 mg/kg b.w./day 1000 mg/kg b.w./day

375 mg/kg b.w./day

500 mg/kg b.w./day

500 mg/kg b.w./day

1125 mg/kg b.w./day

1000 mg/kg b.w./day

500 mg/kg b.w./day

No observable effect level

40,000 ppm (HCT) 1000 ppm (1165 mg/kg b.w./day) (25 mg/kg b.w./day) (based on FC data)

Chronic toxicity/ oncogenecity 18-month dietary (mouse)

2-year dietary (rat) Systematic toxicity 10,000 ppm (HCT) No observable effect level (500 mg/kg b.w./day — males; 640 mg/kg b.w./day — females) (based on FC data) Oncogenicity 10,000 ppm (HCT) No observable effect level (500 mg/kg b.w./day — males; 640 mg/kg b.w./day – females) (based on FC data)

7000 ppm (HCT) 1000 ppm (1200 mg/kg b.w./day) (150 mg/kg b.w./day)

Chapter | 86  Imidazolinones

Chronic toxicity 1-year dietary (dog)

(based on FC data)

Developmental and reproductive toxicity Teratology (rabbit) Maternal No observable effect level Developmental No observable effect level Teratology (rat) Maternal No observable effect level

Developmental

1000 mg/kg b.w./day

(Continued)

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Table 86.2  (Continued) Technical herbicide

Study

No observable effect level Reproduction (multigeneration) (rat) Reproductive Toxicity

Imazapyr

Imazamethabenzmethyl

Imazapic

Imazethapyr

Imazamox

Imazaquin

(HDT) 10,000 ppm (HCT) (800 mg/kg b.w./day — males; 980 mg/kg b.w./day – females) (based on FC data)

(HDT) 4000 ppm (HCT) (320 mg/kg b.w./day)

(HDT) 20,000 ppm (HCT) (1600 mg/kg b.w./day) (based on FC data)

10,000 ppm (HCT) (800 mg/kg b.w./day)

(HDT) 20,000 ppm (HCT) 10,000 ppm (HCT) (1640 mg/kg b.w./day) (800 mg/kg b.w./day) (based on FC data)

Not mutagenic or genotoxic

Not mutagenic or genotoxic

Not mutagenic or genotoxic

Not mutagenic or genotoxic

No observable effect level Gennotoxicity

Categories are EPA Toxicity Categories. b.w., body weight; HDT, highest dose tested; HCT, highest concentration tested; , approximately equal to; FC, foot consumption.

Not mutagenic or genotoxic

Hayes’ Handbook of Pesticide Toxicology

[Gene mutations (Ames Not mutagenic or and mammalian cell); genotoxic in vitro structural chromosomal aberrations; in vivo cytogenetics (mouse micronucleus assay)]

Chapter | 86  Imidazolinones

following 21 days of dermal exposure (6 hours per day, 5 days per week) at the highest doses tested, supporting no observable effect levels (NOELs) at the highest doses tested. Specifically, these no-effect levels were 1000 mg/kg body weight/day for imazapic, imazethapyr, imazamox (28-day study), and imazaquin; and 400 and 200 mg/kg body weight/day for imazapyr and imazamethabenzmethyl, respectively. 90-Day Dietary Toxicity Studies in the Dog Subchronic (90-day) feeding studies were conducted in dogs with imazapyr, imazethapyr, and imazamox. For these subchronic (90-day) feeding studies, the imidazolinone herbicides showed no systemic toxicity at the highest concentrations tested. Specifically, the NOELs were 10,000 ppm for imazapyr and imazethapyr and 40,000 ppm for imazamox. These highest dietary concentrations are equivalent to approximately 250 mg/kg body weight/day for imazapyr and imazethapyr, and equivalent to an approximate daily intake value of 1370 mg/kg body weight/day for imazamox, as calculated from food consumption data. 90-Day Dietary Toxicity Studies in the Rat  For the subchronic (90-day) feeding studies in rats, most of the imidazolinone herbicides showed no systemic toxicity when tested at very high dietary concentrations of 20,000 ppm (imazapyr, imazapic, and imazamox) or 10,000 ppm (imazethapyr and imazaquin). Moreover, these highest dietary concentrations are equivalent to approximate daily intake values of 1740 mg/kg body weight/day (imazapyr), 1625 mg/kg body weight/day (imazapic), 1660 mg/kg body weight/day (imazamox), 820 mg/kg body weight/day (imazethapyr), and 830 mg/kg body weight/day (imazaquin), as calculated from food consumption data. Only imazamethabenz-methyl induced several mild treatment-related effects following 90 days of dietary exposure at the two highest concentrations tested, 5000 and 10,000 ppm. Specifically, slight but consistent decreases in mean body weight occurred for males and females at 5000 and 10,000 ppm, as compared to controls. These reductions in body weights resulted in decreases in overall body weight gain of 8% and 6% for males at 5000 and 10,000 ppm, respectively. For females at both 5000 and 10,000 ppm, overall body weights were decreased by 5%, as compared to controls. No effects on mean body weight or body weight gain were noted for males or females at 1000 ppm, the lowest dietary concentration tested. In addition, statistically significant increased relative (to body weight) liver weights were observed for males at 10,000 ppm, as compared to controls. Increased incidences of hepatocellular hypertrophy were noted for male rats at 5000 ppm (14/19) and 10,000 ppm (20/20), as compared to controls (0/20). The microscopic change of hepatocellular hypertrophy may represent an adaptive response in the liver, that is, induction of microsomal enzymes, which is generally not considered to be a toxic effect (Popp and Cattley, 1991). It is known that this morphological change

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and associated enzyme induction are reversible following withdrawal of chemical treatment (Popp and Cattley, 1991). In conclusion, for the 90-day dietary toxicity study in rats with imazamethabenz-methyl, the no-effect level was the lowest concentration tested, 1000 ppm (equivalent to approximately 87.5 mg/kg body weight/day, as based on actual food consumption data).

86.2.1.3  Chronic Toxicity Studies 1-Year Dietary Toxicity Study in the Dog For the 1year chronic dog studies, imazapyr and imazamox demonstrated no treatment-related effects following dietary exposure at the highest concentration tested (HCT), namely: 10,000 ppm (equivalent to approximately 250 mg/ kg body weight/day) for imazapyr and 40,000 ppm (approximately 1165 mg/kg body weight/day, as based on food consumption data) for imazamox (see Table 86.2). For the other imidazolinone herbicides, imazethapyr, imazapic, imazaquin, and imazamethabenz-methyl, slight treatment-related effects were observed in the respective 1-year chronic dog studies. Specifically, for the 1-year chronic feeding study in dogs with imazethapyr using dietary concentrations of 1000, 5000, and 10,000 ppm, the only treatment-related effects were indicative of a slight anemia, that is, decreased red cell parameters [statistically significant decreases in hematocrit, hemoglobin, red blood cell (RBC) count, mean corpuscular hemoglobin (MCH), mean corpuscular volume (MCV), and mean corpuscular hemoglobin concentration (MCHC)], which were observed at weeks 26 and 52 for females at the mid-concentration (5000 ppm) and the high concentration (10,000 ppm). For this chronic dog study with imazethapyr, no treatment-related histopathological lesions were observed at any dietary concentration, including the highest concentration tested (10,000 ppm). The no observable effect level for the study was 1000 ppm (equivalent to approximately 25 mg/kg/day). Similar effects indicative of anemia were observed in the 1-year chronic dietary dog study with imazapic. In this study, the anemia occurred in both sexes and at higher dietary concentrations of 20,000 and 40,000 ppm. At the highest concentration tested (40,000 ppm), decreased red blood cell parameters (statistically significant decreases in hematocrit, hemoglobin, RBC count, MCH, MCV, and MCHC) were observed at weeks 5, 6, 13, 26, and 52 for males and females with accompanying statistically significant increased numbers of normoblasts and reticulocytes, as compared to controls. In addition, in the 40,000 ppm treatment group, increased incidences of erythropoiesis were observed microscopically in the bone marrow (5 of 6 males, 6 of 6 females) and spleen (4 of 6 males, 2 of 6 females), as compared to controls (0 of 6 males, 0 of 6 females for both bone marrow and spleen). For male and

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female dogs at 20,000 ppm, only transient statistically significant decreases in red blood cell parameters were observed at sporadic time points throughout the study. At the terminal sacrifice, generally slightly increased incidences of erythropoiesis were diagnosed microscopically in the bone marrow of both sexes (2 of 6 males, 1 of 6 females) and spleen of males (1 of 6 males, 0 of 6 females), as compared to controls (0 of 6 males, 0 of 6 females). For male and female dogs at 5000 ppm, no treatment-related effects indicative of anemia were observed. Further, for most dogs (4 of 5 males and 5 of 6 females) at the 40,000-ppm concentration in the chronic dog study with imazapic, slight-to-moderate skeletal muscle myopathy was diagnosed microscopically at the terminal sacrifice of 52 weeks, which was preceded by transient increases (beginning at week 5) in the blood serum of the following enzymes contained in skeletal muscle: creatine kinase, aspartate amino-transferase, and lactate dehydrogenase. In addition, one male dog at 40,000 ppm, that died during the study (week 33) with nontreatment-related bronchopneumonia, showed moderate-to-marked skeletal muscle degeneration on microscopic examination. In contrast, only a limited presence of skeletal myopathy of minimal severity was diagnosed at 20,000 ppm (5 of 6 males and 2 of 6 females) and at 5000 ppm (5 of 6 males and 1 of 6 females). The skeletal myopathy observed at both 5000 and 20,000 ppm was not considered to be adverse because the limited presence of minimal skeletal myopathy at both concentrations was evidenced by only a few fibers out of hundreds evaluated per section per animal. Furthermore, these focal myopathies of minimal severity were not consistently diagnosed in all skeletal muscle sites examined per dog (i.e., vastus and abdominal muscles, diaphragm, and esophagus). Moreover, none of the dogs on study, including the 6 males and 6 females at 40,000 ppm, showed any clinical observations during the 1-year study to indicate any muscle dysfunction. Finally, at the 5000- and 20,000-ppm dietary concentrations, no statistically or biologically significant elevations occurred during the course of the 1-year study in serum enzymes that are normally increased with skeletal muscle myopathy (i.e., creatine kinase, aspartate aminotransferase, or lactate dehydrogenase). For the reasons given previously, the minimal myopathy diagnosed histologically at 5000 and 20,000 ppm was not considered to impair or adversely affect the functional capacity of the affected skeletal muscles. Thus, based on the slight anemia observed at the mid-concentration (20,000 ppm), the lowest dietary concentration (5000 ppm) was regarded as the no observable adverse effect level (NOAEL), equivalent to 135 mg/kg body weight/day, as calculated from actual food consumption data. Similar skeletal muscle myopathies were observed for another imidazolinone herbicide, imazaquin, having similar structure–activity relationships to imazapic and imazethapyr, as mentioned in Section 86.1. For the 1-year chronic feeding

Hayes’ Handbook of Pesticide Toxicology

study in dogs with imazaquin, the dietary concentration of 5000 ppm (highest concentration tested) induced slight anemia as evidenced by bone marrow hyperplasia (erythropoiesis), observed microscopically in 2 of 8 males and 4 of 8 females, and slight skeletal myopathy, observed microscopically in 7 of 8 males and 3 of 8 females. In addition, decreased red blood cell parameters indicative of anemia (statistically significant decreases in hematocrit, hemoglobin, RBC count, MCH, MCV, and MCHC) were observed at weeks 13, 26, and 52 for males and females. Additionally, at 5000 ppm, clinical chemistry parameters indicative of slight skeletal myopathy included statistically significant increased mean serum enzyme levels of creatine kinase, aspartate aminotransferase, and lactate dehydrogenase at weeks 13, 26, and/or 52 for males and females. The NOEL for systemic toxicity for this chronic dog study with imazaquin was the mid-concentration of 1000 ppm (equivalent to approximately 25 mg/kg body weight/day), which is the same NOEL in the 1-year dog study with imazethapyr. Finally, for the 1-year chronic toxicity study in dogs with imazamethabenz-methyl, males and females were fed dietary concentrations of 0, 250, 1000, or 4000 ppm. Body weights at the highest concentration were consistently, but not statistically, lower than the controls for males throughout the study and for females from weeks 1–28. This reduction in body weight resulted in an 11% decrease in overall body weight gain for males at 4000 ppm. There were no treatment-related effects on hematology or clinical chemistry parameters and no treatment-related gross necropsy or microscopic findings. Slight but statistically significant increases in absolute and relative (to body weight) liver weights were noted for females at 4000 ppm, as compared to controls. However, these slight increases in absolute and relative (to body weight) liver weights for females at 4000 ppm are considered to be equivocal because of the absence of statistically significant increases in relative (to brain) liver weights for females at 4000 ppm and the lack of treatment-related histopathological findings in the liver of males or females at 4000 ppm. The absence of hepatocellular hypertrophy in the liver of dogs at 4000 ppm suggests that the small increase in liver weight may not be treatment related. Based on the slight decreases in mean body weights and an 11% decrease in overall body weight gain for males at 4000 ppm, the no observable effect level was 1000 ppm (equivalent to approximately 25 mg/kg body weight/day).

86.2.1.4  Chronic Toxicity/Oncogenicity Studies 18-Month Chronic Toxicity/Oncogenicity Studies in the Mouse For the 18-month mouse feeding studies, all six of the imidazolinone herbicides showed no evidence of potential oncogenicity at the highest dietary concentrations of technical materials tested, namely, 10,000 ppm (imazapyr and imazethapyr), 7000 ppm (imazapic and imazamox), 4000 ppm (imazaquin), and 2100 ppm (imazamethabenz-methyl).

Chapter | 86  Imidazolinones

Furthermore, three of the imidazolinone herbicides showed no systemic toxicity at the highest dietary concentrations tested, namely, imazapyr (at 10,000 ppm) and both imazapic and imazamox (at 7000 ppm). The other three imidazolinones showed mild systemic toxicity in the 18-month mouse studies. Specifically, imazamethabenz-methyl was tested in the mouse at dietary concentrations of 0, 130, 525, and 2100 ppm. Mean body weights for females at 2100 ppm were statistically significantly lower than the controls during the first 8 weeks of the study, which resulted in a 7% decrease in overall body weight gain. At terminal sacrifice, mean absolute and relative (to body weight) thyroid/parathyroid weights were slightly but statistically significantly increased for females at 525 ppm and for both males and females at 2100 ppm, as compared to controls. In addition, mean absolute and relative (to body weight) adrenal gland weights were slightly but statistically significantly increased for males at 2100 ppm. However, these slight organ weight changes are not considered to be adverse because no correlating histopathological changes were noted in the adrenal, thyroid, or parathyroid glands at any dietary concentration in the study. Based on decreased mean body weights and body weight gain for females at 2100 ppm, the no observable adverse effect level (NOAEL) for systemic toxicity was 525 ppm (equivalent to approximately 79 mg/kg body weight/day). Imazethapyr was tested in the mouse for 18 months at dietary concentrations of 0, 1000, 5000, and 10,000 ppm. The only treatment-related effect occurred at the highest concentration tested (10,000 ppm); that is, decreased overall mean body weight gain was noted in both sexes (14% for males and 24% for females) at 10,000 ppm, as compared to controls. Therefore, the NOEL for systemic toxicity was 5000 ppm (equivalent to approximately 750 mg/kg body weight/day). Lastly, imazaquin was tested in the mouse for 18 months at dietary concentrations of 0, 250, 1000, and 4000 ppm. The only treatment-related effect occurred at the highest concentration tested (4000 ppm); that is, decreased mean body weights were noted at 4000 ppm, as compared to controls. Specifically, mean body weights of males at 4000 ppm were statistically significantly decreased during the first 12 weeks of the study, resulting in an 8% decrease in body weight gain for weeks 1–12, as compared to control males. In addition, mean body weights of females at 4000 ppm were statistically significantly lower than controls throughout the study, resulting in a decrease in overall mean body weight gain of 15%, as compared to control females. Therefore, the NOEL for systemic toxicity was 1000 ppm (equivalent to approximately 150 mg/kg body weight/day). 2-Year Chronic Toxicity/Oncogenicity Studies in the Rat In chronic studies performed at high dietary concentrations for 2 years in the rat, all six of the imidazolinone herbicides showed no evidence of potential oncogenicity/carcinogenicity. Specifically, no treatment-related increased incidences of benign or malignant tumors were induced by

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the imidazolinones at the highest concentrations tested, namely, 20,000 ppm (imazapic and imazamox), 10,000 ppm (imazapyr, imazethapyr, and imazaquin), and 4000 ppm (imazamethabenz-methyl). Furthermore, five of the six imidazolinone herbicides showed no systemic toxicity at the highest dietary concentrations tested, namely, imazapic and imazamox (at 20,000 ppm) and imazapyr, imazethapyr, and imazaquin (at 10,000 ppm). Only imazamethabenz-methyl showed mild systemic toxicity in the 2-year rat feeding study. Specifically, imazamethabenz-methyl was tested in the rat at dietary concentrations of 0, 250,1000, and 4000 ppm. Mean body weights were slightly but consistently decreased for males and females at 4000 ppm, and for females at 1000 ppm, as compared to controls. For males at 4000 ppm a decreased overall mean body weight gain of 6% was noted, as compared to controls. For females at 1000 and 4000 ppm, respective decreased overall mean body weight gains of 6% and 11% were noted, as compared to controls. There were no treatment-related microscopic lesions with the exception of a statistically significant increased incidence of thymic epithelial hyperplasia in females at 4000 ppm (21/53), as compared to controls (7/49). However, for this 2-year (lifetime) rat study with imazamethabenz-methyl, the thymic hyperplasia did not progress to a neoplastic lesion nor has such a progression been described in the literature. Therefore, for this reason, this hyperplastic change in females at 4000 ppm appears to have only an equivocal toxicologic significance. In conclusion, based on the results from this study, the NOEL for systemic toxicity was 250 ppm (equivalent to approximately 12.5 mg/kg body weight/day). For imazamethabenz-methyl, the increased incidence of hepatocellular hypertrophy in the liver of male rats at both 5000 and 10,000 ppm, which was noted at termination of the 90-day dietary toxicity study, was not observed at either the 12-month interim sacrifice or the terminal sacrifice of the 2-year chronic rat study at dietary concentrations up to and including 4000 ppm. The absence of this microscopic finding indicates that adaptation probably occurred in the liver following prolonged exposure. In conclusion, the collective results from the oncogenicity studies in both the rat and the mouse indicate a lack of oncogenic/carcinogenic potential for all six imidazolinone herbicides.

86.2.2  Absorption, Distribution, Metabolism, and Excretion In the rat metabolism studies with five of the six imidazolinone herbicides, only minimal metabolism is demonstrated. Following single oral gavage doses of imazapyr, imazapic, imazethapyr, imazamox, or imazaquin, these imidazolinone herbicides are rapidly absorbed and excreted, as evidenced by the presence of greater than 70% of unchanged parent

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compound in the urine within 24–48 h. Although imazamethabenz-methyl is also rapidly absorbed in the rat, greater than 60% becomes metabolized via hydrolysis of the ester to imazamethabenz acid, which is rapidly excreted in the urine within 24 h. The presence of a significant amount of unchanged parent or metabolite in the urine within 24–48 h, indicates rapid absorption of the imidazolinone herbicides from the gastrointestinal tract following a single oral gavage dose, as well as a low potential for bioaccumulation of parent compound or acid metabolite in mammalian tissues.

86.2.3  Effects on Organs and Tissues The only consistent treatment-related effects that were observed in the extensive toxicological profile of the imidazolinone herbicides were slight-to-moderate skeletal myopathy and/or slight anemia in dogs, occurring in the 1-year dietary toxicity studies with three structurally similar imidazolinones (imazapic, imazaquin, and imazethapyr). Specifically, both of these treatment-related effects were seen in male and female dogs treated with imazapic. Slight anemia at 20,000 ppm (mid-concentration) and 40,000 ppm (highest concentration tested) was noted by decreases in hematological parameters and by histopathological examination, whereas slight-tomoderate skeletal myopathy at 40,000 ppm was determined by clinical chemistry evaluation and histopathological examination. Similarly, both of these treatment-related effects were seen in male and female dogs treated with imazaquin. Slight anemia at 5000 ppm (highest concentration tested) was noted by decreases in hematological parameters and by histopathological examination, whereas slight skeletal myopathy at 5000 ppm was determined by clinical chemistry evaluation and histopathological examination. In contrast, only slight anemia was observed by hematological parameters in female dogs with imazethapyr at both 5000 ppm and 10,000 ppm (highest concentration tested). There was no evidence of potential carcinogenicity in gross necropsy observations or in microscopic examinations of the full battery of tissues in the rat or mouse for any of the imidazolinone herbicides.

86.2.4  Effects on Reproduction Results from the reproductive and developmental toxicity studies indicate that the six imidazolinone herbicides are not reproductive toxicants, developmental toxicants, or teratogens (Table 86.2). Specifically, in the multigeneration reproductive toxicity studies conducted in rats, the imidazolinone herbicides did not affect reproductive performance, nor was there evidence of any significant prenatal or postnatal effects. All six reproduction studies support reproductive NOELs at the highest concentrations tested, namely, 20,000 ppm (imazapic and imazamox), 10,000 ppm (imazapyr, imazethapyr, and imazaquin), or 4000 ppm

Hayes’ Handbook of Pesticide Toxicology

(imazamethabenz-methyl). These highest dietary concentrations are equivalent to approximate daily intake values of 1600 mg/kg body weight/day (imazapic), 1640 mg/kg body weight/day (imazamox), and 800 mg/kg body weight/day for males and 980 mg/kg body weight/day for females (imazapyr), as calculated from food consumption data. The preceding doses are equivalent to approximately 800 mg/kg body weight/day (imazethapyr and imazaquin) or 320 mg/kg body weight/day (imazamethabenz-methyl). Further, the teratology studies, which evaluated potential developmental toxicity of the six imidazolinone herbicides in rabbits and rats, revealed no evidence of developmental toxicity or teratogenic effects for fetuses of either species. All six imidazolinones, as tested in rabbits and rats, showed developmental NOELs equal to or higher than the maternal NOEL/ NOAELs. For the rabbit teratology studies, the developmental NOELs were 1000 mg/kg body weight/day, the highest dose tested (HDT) for imazethapyr; 900 mg/kg body weight/ day (HDT) for imazamox; 700 mg/kg body weight/day (HDT) for imazapic; 500 mg/kg body weight/day for imazamethabenz-methyl, 500 mg/kg body weight/day (HDT) for imazaquin; and 400 mg/kg body weight/day (HDT) for imazapyr. For these rabbit studies, the maternal NOELs were either the same dose level as the developmental no-effect level (e.g., 500 mg/kg body weight/day for imazamethabenzmethyl; 400 mg/kg body weight/day for imazapyr) or lower dose levels of 500 mg/kg body weight/day for imazapic, 300 mg/kg body weight/day for imazethapyr and imazamox, or 250 mg/kg body weight/day for imazaquin. Importantly, for all six imidazolinone herbicides, no treatment-related teratogenic effects were observed in the rabbit fetuses. For the rat teratology studies, the developmental NOELs were 1000 mg/kg body weight/day, the HDT for imazapyr, imazamethabenz-methyl, imazapic, and imazamox; 500 mg/kg body weight/day (imazaquin); and 375 mg/kg body weight/day (imazethapyr). For these rat studies, the maternal NOEL/ NOAELs were either the same dose level as the developmental no effect level (e.g., 1000 mg/kg body weight/day for imazamethabenz-methyl and imazapic; 500 mg/kg body weight/day for imazaquin) or lower dose levels of 500 mg/kg body weight/day for imazamox, 375 mg/kg body weight/day for imazethapyr or 300 mg/kg body weight/day for imazapyr. Importantly, for all six imidazolinone herbicides, no treatmentrelated teratogenic effects were observed in the rat fetuses. In conclusion, based on the results given previously for the multigeneration reproduction studies and the rabbit and rat teratology studies, the imidazolinone herbicides demonstrate a lack of reproductive toxicity and are neither selective developmental toxicants nor teratogens in either the rabbit or the rat.

86.2.5  Pathology The only consistent treatment-related effects that were observed microscopically in the extensive toxicological

Chapter | 86  Imidazolinones

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profile of the imidazolinone herbicides were slight anemia and slight-to-moderate skeletal myopathy in dogs, occurring in the 1-year dietary toxicity studies with two structurally similar imidazolinones (imazapic and imazaquin). Specifically, both of these treatment-related effects were seen in male and female dogs treated with imazapic. Slight anemia was noted by histopathological examination (erythropoiesis in the bone marrow and spleen) at 20,000 ppm (mid-concentration) and 40,000 ppm (highest concentration tested). Slight-to-moderate skeletal myopathy was also diagnosed by histopathological examination (skeletal muscle degeneration) at 40,000 ppm (terminal sacrifice). Similarly, both of these treatment-related effects were seen in male and female dogs treated with imazaquin. Slight anemia (erythropoiesis in the bone marrow) was diagnosed at 5000 ppm (highest concentration tested). In addition, slight skeletal myopathy (skeletal muscle degeneration) was observed at 5000 ppm by histopathological examination. There was no evidence of potential carcinogenicity for any of the imidazolinone herbicides from evaluations of the gross necropsy observations or the microscopic findings of the full battery of tissues in the rat or mouse.

process of the imidazolinone herbicides or for workers involved in mixing/loading/applying the end-use products of the imidazolinone herbicides for crop or noncrop uses. Results from the acute dermal and oral toxicity data, as cited in Section 86.2, indicate that the imidazolinone herbicides do not pose any acute dermal or dietary risks. Furthermore, because of their relatively low toxicity to mammals, the imidazolinone herbicides demonstrated relatively high no observable effect levels for potential systemic toxicity in the long-term studies, as cited in Section 86.2. In the absence of genotoxic, carcinogenic, reproductive, or teratogenic effects, a safety factor of 100 (10 for interspecies differences, 10 for intraspecies differences) is appropriate to calculate the acceptable daily intake (ADI) for chronic human exposure. Applying a low (100fold) safety factor to these relatively high systemic NOELs from the long-term studies with the imidazolinone herbicides results in ADIs that are relatively high. These relatively high ADIs suggest that this class of compounds does not pose a concern for chronic dietary exposure to humans.

86.2.6  Genotoxicity Studies

There are no known cases of accidental or deliberate poisonings in humans. Because of their low toxicity profile and rapid excretion rate (see preceding discussion), the development of any physiological antidotes for the imidazolinone herbicides appears unnecessary. For formulated products, it is advisable to consult the specific material safety data sheet (MSDS) for emergency and first-aid procedures.

As presented in Table 86.2, based on the battery of in vitro and in vivo assays, the imidazolinone herbicides show a lack of potential genotoxic activity. This series of tests comprised the genotoxicity testing data requirements for all three categories [i.e., gene mutations (Ames and mammalian cell), in vitro structural chromosomal aberrations, and in vivo abnormal cytogenetics such as detected in the mouse micronucleus assay using bone marrow cells].

86.3  Toxicity to humans 86.3.1  Use Experience There are no incident reports attributable to the active ingredients for workers involved in the manufacturing

86.3.2  Treatment of Poisoning

Reference Popp, J. A., and Cattley, R. C. (1991). Hepatobiliary system. In “Handbook of Toxicologic Pathology” (W. M. Haschek and C. G. Rousseaux, eds.), pp. 279–314. Academic Press, San Diego.

Chapter 87

Toxicology of Triazolopyrimidine Herbicides Richard Billington, Sean C. Gehen Dow AgroSciences LLC, Indianapolis, Indiana

Thomas R. Hanley Jr. Syngenta Crop Protection, Inc., Greensboro, North Carolina

87.1  Introduction The triazolopyrimidines are herbicides used for the preemergent and postemergent control of broadleaf weeds in a variety of crops. The general structure of this class of chemistry is a substituted triazolopyrimidine connected to a substituted phenyl (cloransulam-methyl, diclosulam, florasulam, flumetsulam, metosulam, penoxsulam) or pyridine (pyroxsulam) ring through a sulfonamide bridge, as presented in Figure 87.1. The substituents of the various members of this class are presented in Table 87.1. The herbi­cidal mode of action is through inhibition of acetolactate synthase (ALS) in plants, though the mechanism appears to be different from that of sulfonylureas. Acetolactate synthase (EC 4.13.18), also known as acetohydroxyacid synthase, is a key enzyme in the synthesis of

the ­ branched-chain aliphatic amino acids leucine, isoleucine, and valine. Inhibition of this enzyme in plants results in cessation of cell growth and division, leading to the death of suscep­tible plants. However, this enzyme is lacking in humans and other animals, which accounts for the low mammalian toxicity of these chemicals. R1

R4

R2 N 4

N

1

5

N

3

R3

2

R5

R6

Figure 87.1  Generic structure of the triazolopyrimidine herbicides.

Table 87.1  Substituents of Triazolopyrimidine Sulfonamide Herbicides 1

2

3

4

5

R1

R2

R3

R4

R5

R6

(1,5c) Clorasulam-methyl

N

C

C

NH

SO2

H

CO2CH3

Cl

OCH2CH3

F

H

Diclosulam

N

C

C

NH

SO2

H

Cl

Cl

OCH2CH3

F

H

Florasulam

N

C

C

NH

SO2

H

F

F

OCH3

H

F

Penoxsulama

N

C

C

SO2

NH

H

OCH2CHF2

CF3

OCH3

H

OCH3

(1,5a) Flumetsulam

C

N

C

NH

SO2

H

F

F

H

CH3

—

Metosulam

C

N

C

NH

SO2

CH3

Cl

Cl

OCH3

OCH3

—

C

N

N

SO2

NH

H

CF3

OCH3

OCH3

OCH3

—

a

Pyroxsulam a

Sulfonamide linkage reversed

Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

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Comprehensive toxicological testing has been conducted with all of these substances according to national and international regulatory frameworks for the approval and registration of pesticides and their products. The methods used for the studies summarized in this review typically comply with the regulatory test guidelines – as published by the Organization for Economic Cooperation and Development (OECD), the U.S. Environmental Protection Agency (U.S. EPA), the European Union (EU), and Japan Ministry of Agriculture, Forestry, and Fisheries (JMAFF) to determine potential health effects – and Good Laboratory Practice (GLP) regulations. Studies typ­ ically involve oral administration of the test substance by incorporation into the animal feed, which can be achieved either using a fixed dietary concentration throughout the course of the study or by regularly varying the concentration of the substance in the feed to achieve a constant dosage in terms of milligrams per kilogram of body weight per day. The notable exceptions to this typical approach are acute oral studies, including neurotoxicity, developmental toxicity studies, and in vivo genotoxicity studies, which typically involve oral administration of the substance suspended in an aqueous vehicle by gavage. These studies are used to characterize any potential hazards resulting from acute, subchronic, or chronic exposure and to derive reference doses for humans (e.g., an Acceptable Daily Intake, ADI). These are required to enable the acceptability of predicted exposures to the triazolopyrimidines under all anticipated use scenarios to be determined as part of the comprehensive human health-based risk assessments. In general, triazolopyrimidine herbicides have very low mammalian toxicity as determined by acute, shortterm, long-term (chronic), genotoxicity, reproduction, developmental, and neurotoxicity tests. The kidneys and/or liver are the primary organs affected by repeated exposure, and, even then, in most cases the effects represent adaptive responses to very high levels of exposure. From a pharmacokinetic perspective, oral absorption is rapid, as is excretion, with no evidence of accumulation. In all cases, the sulphonamide bridge remains intact and the parent molecule is the principle component in excreta, mainly in urine. Although metabolism is limited, in some cases many metabolites are produced, which typically comprise ring demethylation and hydroxylation with or without formation of various types of conjugations. The mammalian toxicity of these materials is reviewed in this chapter.

87.2  Cloransulam-methyl 87.2.1  Identity, Properties, and Uses 87.2.1.1  Chemical Name The IUPAC name for cloransulam-methyl is methyl 3-chloro-2-(5-ethoxy-7-fluoro[1,2,4]triazolo[1,5-c]pyrimidin2-ylsulfonamido)benzoate; the CAS name is methyl

3-chloro-2-[[(5-ethoxy-7-fluoro[1,2,4]triazolo[1,5-c] pyrimidin-2-yl)sulfonyl]amino]benzoate.

87.2.1.2  Structure See Figure 87.1 and Table 87.1.

87.2.1.3  Synonyms Cloransulam-methyl is also known as XR-565, or XDE565, and is sold as FirstRate herbicide in the United States, and as Pacto and Supra herbicides in South America. (All trade names used in this chapter are registered trademarks of Dow AgroSciences LLC.) The Chemical Abstract Service (CAS) registry number is 147150-35-4.

87.2.1.4  Physical and Chemical Properties The empirical formula for cloransulam-methyl is C15H13 ClFN5O5S, with a molecular weight of 429.8. It is a solid at room temperature, with a low vapor pressure (3   1016 mmHg at 25°C). The water solubility is pH-dependent, with values of 2.96 mg/l at pH 5, 184 mg/l at pH 7, and 3430 mg/l at pH 9 (20°C). The log Kow is estimated at 3.7; the pKa is 4.81.

87.2.1.5  Uses Cloransulam-methyl is used as a soil-applied or incorporated preemergence or postemergence broadleaf herbicide in soybeans at maximum label rates of 44 g per hectare soil applied and 18 g per hectare postemergence.

87.2.2  Toxicity to Laboratory Animals 87.2.2.1  Acute Exposure The acute toxicity of cloransulam-methyl is low. The acute oral LD50 in the rat was greater than 5000 mg/kg body weight (bw) in both males and females and the dermal LD50 in the rabbit was greater than 2000 mg/kg bw. The 4-h inhalation LC50 in the rat was greater than 3.77 mg/l of air, which was the highest attainable respirable aerosol concentration. Cloransulam-methyl produced no indications of dermal irritation in rabbits or delayed contact skin sensitization in the guinea pig, and only slight transient eye irritation in the rabbit following acute exposure (Bradley et al., 1992; Cosse and Berdasco, 1992a,b,c,d; U.S. EPA, 1997a,b).

87.2.2.2  Repeated Exposure Cloransulam-methyl was evaluated in subacute and subchronic dietary studies in rats, mice, and dogs. The primary target organs identified in these studies were the kidney (rat and mouse), the liver (rat, mouse, and dog), and the thyroid (rat).

Chapter | 87  Toxicology of Triazolopyrimidine Herbicides

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In the Fischer 344 (F344) rat, dosages of 100–1000 mg/ kg/day for 2 weeks produced slight decreases in red blood cell parameters and urine-specific gravity in males, and slightly increased cecal and liver weights in females at 1000 mg/kg/day. The no-observed-effect level (NOEL) was 500 mg/kg/day. Dosages of 100–1000 mg/kg/day for 13 weeks produced treatment-related kidney changes comprising very slight to moderate hypertrophy of collecting tubule epithelial cells and/or slight vacuolation of the renal proximal tubular epithelium consistent with fatty changes in all treated groups. Decreased body weight gain and feed consumption, very slight hepatocellular vacuolation, and slight thyroid follicular hypertrophy were also seen at 500 and 1000 mg/kg/day (Haut et al., 1991, 1992a; Stebbins and Haut, 1994). Dosages of 100, 500, or 1000 mg/kg/day fed to B6C3F1 mice for 2 weeks produced slight hepatocellular hyper­trophy at 500 mg/kg/day and above in males, and at 1000 mg/kg/day in females. The NOEL was 100 mg/kg/day. Dosages ranging from 50 to 1000 mg/kg/day given for 13 weeks produced slight centrilobular and midzonal hepatocellular hypertrophy at 100 mg/kg/day and above in males, and at 500 mg/kg/day and above in females. Electron microscopy characterized the hypertrophy as an increase in rough endoplasmic reticulum (RER) with a decrease in cytoplasmic glycogen content. Kidney effects in mice consisted of decreased vacuolation of the renal tubules, consistent with decreased cytoplasmic lipid, accompanied by lower kidney weights at 500 mg/kg/day and higher. The subchronic lowest observed effect level (LOEL) and NOEL values in mice were 100 and 50 mg/kg/day, respectively (Haut et al., 1992b; Stebbins and Haut, 1993). Cloransulam-methyl, when fed to dogs for 2 weeks at dosages of 500 mg/kg/day or higher, produced hepatic inflammation, degeneration, and necrosis. No effects were seen at dosages of 200 mg/kg/day or lower. In a subchronic study, dogs exhibited a taste aversion to this material at dosages of 200 mg/kg/day and above, which resulted in a combination of impaired nutritional status and toxicity of the material. A dosage of 40 mg/kg/day resulted in lower body weight gains. Histologic examination did not identify a target organ, although a subsequent chronic study in dogs identified the liver as the primary target organ. Based on decreased body weights, a subchronic NOEL was not established in dogs (Stebbins et al., 1996; Szabo et al., 1992). In a 21-day repeated dermal application study in rabbits, cloransulam-methyl at dosages of 100, 500, or 1000 mg/kg/day produced slight anemia in female rabbits at the highest dosage. Male rabbits were unaffected at 1000 mg/kg/day and the NOEL in females was 500 mg/kg/ day (Gilbert and Yano, 1995a).

344 rats were fed cloransulam-methyl at dosages of 10– 325 mg/kg/day. Body weight gain was decreased at the highest dosage. Treatment-related histologic effects were limited to the kidneys and thyroid. Hypertrophy of a population of renal collecting duct epithelial cells identified as -intercalated cells was reported in males and females fed 325 mg/kg/day. (A similar histologic change noted in rats, mice, and dogs following exposure to florasulam will be discussed later in this chapter.) Vacuolation of the proximal tubules (consistent with fatty changes) in males fed cloransulam-methyl at 325 mg/kg/day, and females fed 75 or 325 mg/kg/day, and an increase in the incidence of mineralization of the renal pelvis in males fed 75 or 325 mg/kg/day also were present. Thyroid changes were confined to the high-dosage males (325 mg/kg/day) and consisted of hyperplasia and hypertrophy of follicular cell epithelium. The NOEL in this study was 10 mg/kg/day (Jeffries et al., 1995a). B6C3F1 mice were fed diets containing cloransulammethyl at dosages of 10–1000 mg/kg/day for 2 years. As was seen in mouse subchronic studies, the liver was the primary target organ, with effects also noted in the kidneys. Increases in liver weights in males at 100 mg/kg/ day and females at 1000 mg/kg/day, and centrilobular hypertrophy in males at 100 mg/kg/day were the only treatment-related effects noted in the liver. Kidney weights were decreased in males at 1000 mg/kg/day and females at 100 mg/kg/day. In the kidneys, depletion of the normal epithelial cytoplasmic vacuoles, and decreases in the incidence of renal mineralization and renal tubular degeneration were noted in males at 100 mg/kg/day. All of these histologic changes were interpreted to be either incidental or adaptive-physiologic responses to the test material rather than adverse toxic effects. The NOEL in mice following chronic exposure was 10 mg/kg/day (Jeffries et al., 1995b). There was no evidence of a tumorigenic or carcinogenic response in either mice or rats following long-term exposure. In a 1-year chronic toxicity study, beagle dogs were fed dosages of 5–50 mg/kg/day. The only treatment-related effects were in the liver and consisted of a slight-to-moderate increase in accumulation of pigment in Kuppfer cells and hepatocytes, and slight centrilobular and midzonal hepatocellular hypertrophy at 10 mg/kg/day, with changes in hepatic-related serum chemistry parameters at 50 mg/kg/day (Szabo and Davis, 1994). The U.S. EPA considered 10 mg/kg/day the NOEL in this study (U.S. EPA, 1997a).

87.2.2.3  Chronic Toxicity and Carcinogenicity

In a battery of tests, cloransulam-methyl showed no evidence of genotoxic potential. These tests included a bacterial reverse mutation assay (Ames test), an in vitro cytogenetic assay in Chinese hamster ovary cells (CHO/HGPRT

The chronic toxicity of cloransulam-methyl has been evaluated in rats, mice, and dogs. In a 2-year study, Fischer

87.2.2.4  Mutagenicity

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assay), an in vitro chromosomal aberration assay in rat lymphocytes, and an in vivo cytogenetic assay in mouse bone marrow cells (U.S. EPA, 1997a, b).

87.2.2.5  Neurotoxicity The neurotoxic potential of cloransulam-methyl was evaluated in specialized studies. No neurotoxicologic effects were observed in rats following acute gavage exposure of up to 2000 mg/kg (highest dose tested). A complete battery of neurologic tests including functional observations (handheld and open field observations, grip strength, and landing foot splay), motor activity, and detailed neurohistopathology was conducted following 13-week exposure via the diet to dosages up to 1000 mg/kg/day. No treatmentrelated neurotoxic effects were observed in any of these measures (Shankar et al., 1993; Spencer et al., 1995).

87.2.2.6  Reproductive Toxicity Cloransulam-methyl had no effect on reproduction or embryofetal development. In a multigeneration reproduction study in Sprague-Dawley (SD) rats, dosages of 100 mg/kg/day and above produced kidney and thyroid effects in the adults consistent with effects seen in subchronic and chronic studies. The NOEL for parental animals was 10 mg/kg/day. No effects on reproductive performance or neonatal survival were seen even at the high dosage of 500 mg/kg/day. In a developmental tox­icity study in CD rats, gavage dosages of up to 1000 mg/kg/day (limit test) on gestation days 6–15 produced no maternal or developmental toxicity. In a developmental toxicity study in New Zealand White rabbits administered gavage dosages of 0, 30, 100, or 300 mg/kg/day on gestation days 7–19, maternal weight gain and feed consumption were affected only at 300 mg/kg/day. No adverse embryonal or fetal effects were noted at any dose level (Vedula et al., 1992; Zablotny et al., 1993, 1994).

87.2.2.7  Absorption, Distribution, Metabolism, and Excretion Metabolism studies were conducted with 14C-radiolabeled cloransulam-methyl in the F344 rat using dose levels of 5 or 1000 mg/kg. At 5 mg/kg, over 90% of either a single dose or repeated (15 days) doses was absorbed. At 1000 mg/kg, only 28–30% of a single dose was absorbed. Urinary elimination was rapid in both cases with half-lives of approximately 6–9 h. A higher percentage of the 5-mg/ kg dose was excreted in the urine by females (68–80%) than by males (40–50%) and these sex-dependent differences in disposition of the 5-mg/kg dose were attributed to more efficient elimination of unchanged cloransulammethyl in the female versus male kidney. Analyses of urine and fecal extracts indicated that parent cloransulam-methyl

Hayes’ Handbook of Pesticide Toxicology

accounted for the majority of the excreted radiolabeled material. The only metabolite present at amounts greater than 5% was identified as the 4-OH phenyl derivative of cloransulam-methyl. Other minor metabolites included a hydroxylation of the pyrimidine ring, though the position of hydroxylation was not identified, and an N-acetyl cysteine conjugate of the parent material. Due to rapid elimination, cloransulam-methyl has little potential to accumulate upon repeated administration (Domoradzki et al., 1995; Nolan et al., 1995).

87.2.3  Toxicity to Humans No studies are available on intentional human exposure. However, the risk to humans from exposure to cloransulam-methyl following normal use patterns is low. No detectable residues were found either in soybeans or, in most cases, in soybean forage or hay at a limit of detection of 0.005 ppm, and accumulation is unlikely based on plant and animal data. Tolerance levels of 0.02 ppm in soybean, 0.1 ppm for soybean forage, and 0.2 ppm for soybean hay have been established (U.S. EPA, 1997a). Using conservative estimates that assume 100% of crops contain the tolerance level, and a reference dose (RfD) of 0.10 mg/kg/ day (based on the NOEL from the chronic dog study), the calculated maximum potential average daily dose from all sources indicates use of 0.2% of the RfD in the subgroup with the highest aggregate exposure (nonnursing infants). The U.S. EPA estimated the margin of exposure (MOE) for occupational exposure to cloransulam-methyl to be between 2500 and 14,000, based on the use of a NOEL of 10 mg/kg/day (U.S. EPA, 1997a).

87.3  Diclosulam 87.3.1  Identity, Properties, and Uses 87.3.1.1  Chemical Name The IUPAC name for diclosulam is 2,6-dichloro-5-ethoxy7-fluoro[1,2,4]triazolo[1,5-c]pyrimidine-2-sulfonanilide; the CAS name is N-(2,6-dichlorophenyl)-5-ethoxy-7-fluor o[1,2,4]triazolo[1,5-c]pyrimidine-2-sulfonamide.

87.3.1.2  Structure See Figure 87.1 and Table 87.1.

87.3.1.3  Synonyms Diclosulam is also known as XR-564 or XDE-564 and is sold primarily in the United States and South America under the trade names including Strongarm, Spider, and Crosser herbicides. The CAS registry number is 145701-21-9.

Chapter | 87  Toxicology of Triazolopyrimidine Herbicides

87.3.1.4  Physical and Chemical Properties The empirical formula of diclosulam is C13H10Cl2FN5O3S, with a molecular weight of 406.2. Diclosulam is a solid, with a burnt vanilla odor, though the vapor pressure is low (5  1015 mmHg at 25°C). The water solubility is pHdependent and increases with increasing pH, from 117 mg/l at pH 5 to 4290 mg/l at pH 9, and the log Kow values are 0.047 at pH 7 and 0.448 at pH 9.

87.3.1.5  Uses Diclosulam is a soil-applied, preplanting broadleaf herbi­ cide for use in soybeans and peanuts at maximum label rates 35 and 26 g/ha, respectively.

87.3.2  Toxicity to Laboratory Animals 87.3.2.1  Acute Exposure The acute toxicity of diclosulam is low. The acute oral LD50 in the rat was greater than 5000 mg/kg, the dermal LD50 in the rabbit was greater than 2000 mg/kg, and the 4-h inhalation LC50 in the rat was greater than 5.0 mg/l of air. Diclosulam produced no indications of dermal irritation in rabbits or sensitization in the guinea pig, and only very slight transient eye irritation in the rabbit following acute exposure (Cieszlak and Clements, 1993; Gilbert, 1993a,b, c,d,e; U.S. EPA, 1998).

87.3.2.2  Repeated Exposure The primary target organs identified in dietary toxicity studies were the kidneys (rat) and the liver (rat, mouse, and dog). In the F344 rat, dosages of 500 and 1000 mg/kg/day for 2 weeks resulted in increased liver weights and enlarged ceca in males with no histopathologic changes. The NOEL was 100 mg/kg/day in males and 1000 mg/kg/day in females. Rats were given dietary dosages of 50–1000 mg/kg/day for 13 weeks. At 500 and 1000 mg/kg/day, body weights were decreased, and kidney and liver weights were increased. Very slight to moderate treatment-related hepatocellular hypertrophy was observed in males at 100 mg/kg/day and above, and in females at 1000 mg/kg/day. Kidney changes characterized as slightly to moderately decreased intracellular protein in the proximal tubule epithelium were seen in male rats at 500 mg/kg/day and above, secondary to slightly lower feed consumption in these animals. Slight decreases in red blood cell parameters were noted at 100 mg/kg/day and above. The NOEL from this study was 50 mg/kg/day (Stewart et al., 1992a; Szabo and Davis, 1993a). Dietary exposure of B6C3F1 mice to dosages of 100– 1000 mg/kg/day for 2 weeks resulted in slightly decreased kidney weights in both males and females at the high-dose level (1000 mg/kg/day) and slightly decreased hepatocellular vacuolation (consistent with decreased glycogen content) in

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females at 500 mg/kg/day and above. The NOEL was 500 mg/ kg/day in males and 100 mg/kg/day in females. Dosages of 100–1000 mg/kg/day were given to B6C3F1 mice for 13 weeks. Significant body weight effects were seen in males at 1000 mg/kg/day and in females at 500 and 1000 mg/kg/day, and slight-to-moderate hepatocellular hypertrophy was the primary histopathologic change noted at 500 and 1000 mg/kg/ day in males and females, respectively. Kidney weights were lower in males and females at 500 mg/kg/day and above, but there were no correlative changes in clinical chemistry or histopathologic parameters. The NOEL for subchronic ex­posure in the mouse was 100 mg/kg/day (Grandjean and Szabo, 1993; Stewart et al., 1993). Beagle dogs were given diclosulam at dosages of 50– 500 mg/kg/day for 2 weeks. Dosages of 250 mg/kg/day and above were unpalatable and resulted in severely decreased weight gain or weight loss, degenerative changes in the kidneys, and hepatocellular necrosis. A dosage of 50 mg/kg/ day produced microfocal hepatocellular necrosis in males, but no effects in females. In a subchronic study, dogs were given dosages of 0, 5, 25, or 100 mg/kg/day for 13 weeks. Slight, diffuse centrilobular hepatocellular hypertrophy was observed at 25 mg/kg/day. Higher dosages proved to be unpalatable, with secondary toxicity associated with inanition superimposed on the effects of diclosulam on the liver. The subchronic NOEL for the dog was 5 mg/kg/day (Swaim and Szabo, 1992; Szabo and Rachunek, 1992). In a 21-day repeated dermal application study in rabbits, no dermal or systemic effects were seen at 1000 mg/kg/day, the highest dosage tested (Redmond and Kociba, 1996).

87.3.2.3  Chronic Toxicity and Carcinogenicity Chronic studies in rodents with diclosulam produced adaptive changes in the kidney as the primary effect. In a 2-year study in Fischer 344 rats, decreased body weight and weight gain were observed at 400 mg/kg/day, along with changes in hematology, clinical chemistry, and urinalysis parameters associated with the decreased body weight. Histologically, a slight alteration in tubular morphology, mostly within the corticomedullary junction, was observed in the kidneys at 100 and 400 mg/kg/day. This subtle change in the cytologic character and architecture was considered a slight alteration of the normal physiologic state, rather than a pathologic effect indicative of a toxic injury. No effects were noted in rats at 5 mg/kg/day (Minnema, 1996a). Chronic dietary exposure of B6C3F1 mice to dosages of 50–500 mg/kg/day diclosulam for 2 years produced no treatment-related effects on survival, body weight, feed consumption, or clinical observations. The primary histologic change noted in male mice was a reduced vacuolation of the kidney tubular epithelium at all dose levels at the interim and terminal sacrifices, which correlated with decreased absolute and relative kidney weights. In female mice, minimal focal dilation with hyperplasia of the lining

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epithelium of renal cortical tubules was seen at 100 mg/kg/ day and above. In males, this same focal dilation was seen spontaneously across all groups, including controls. There appeared to be no biologic or toxicologic significance to these microscopic changes. The no-observed-adverse-effect level (NOAEL) in mice following chronic exposure was 50 mg/kg/day (Minnema, 1996b). There was no evidence of tumorigenicity or carcinogenicity in either mice or rats. In beagle dogs fed dosages of 2–25 mg/kg/day diclosulam for 1 year, only slight elevations in mean alkaline phosphatase and creatinine levels in dogs given 25 mg/kg/ day were observed. These slight elevations, however, were considered reflective of the normal variability in this species, and 25 mg/kg/day was the NOEL (Walker, 1996).

87.3.2.4  Mutagenicity In a battery of tests, diclosulam showed no evidence of genotoxic potential. These tests included a bacterial reverse mutation assay (Ames test), an in vitro cytogenetic assay in Chinese hamster ovary cells (CHO/HGPRT assay), an in vitro chromosomal aberration assay in rat lymphocytes, and an in vivo cytogenetic assay in mouse bone marrow (U.S. EPA, 1998).

87.3.2.5  Neurotoxicity No neurotoxicologic effects were noted in rats following acute gavage exposure to up to the limit dose of 2000 mg/kg or in a complete battery of neurologic tests including detailed histopathologic examination following 1 year of exposure via the diet to dosages up to 400 mg/kg/day (Mattsson et al., 1996; Minnema, 1996c).

87.3.2.6  Reproductive Toxicity Treatment with diclosulam had no effect on reproduction or fetal development. In a multigeneration reproduction study in Sprague-Dawley rats at dietary dosages up to 1000 mg/ kg/day, no indications of parental or reproductive toxicity were seen. Gavage dosages of up to 1000 mg/kg/day to pregnant Sprague-Dawley rats on gestation days 6–15 produced no maternal or developmental toxicity. In New Zealand White rabbits, no developmental effects were noted even at gavage dosages up to 650 mg/kg/day on gestation days 7–19, which severely affected maternal feed consumption and weight gain. The maternal NOEL in rabbits was 65 mg/ kg/day, whereas the developmental NOEL was 650 mg/kg/ day (Morseth, 1994; Zablotny, 1996; Zablotny et al., 1996).

87.3.2.7  Absorption, Distribution, Metabolism, and Excretion Metabolism studies conducted with 14C-diclosulam in the F344 rat using dose levels of 5 or 500 mg/kg revealed that

Hayes’ Handbook of Pesticide Toxicology

approximately 80% of a single or repeated (15 days) low dose was absorbed by both males and females. At 500 mg/kg, only 15–20% of a single dose was absorbed. Urinary elimination was rapid in both cases with half-lives of approximately 7–12 h. A higher percentage of the 5-mg/kg dose was excreted in the urine by females (62–68%) than by males (39–43%), with the remainder of the absorbed dose eliminated in the feces. At 500 mg/kg, the majority of the administered dose (82–85%) was found in the feces, with only 6–12% eliminated via the urine in both males and females. Within 72 h, less than 3% of the dose remained in the tissues and carcass in all dose groups. The primary urinary and fecal excretion products were identified as unchanged diclosulam and an OH-phenyl oxidation product. In addition, the N-acetyl cysteine conjugate of diclosulam, and the S-oxide of the N-acetyl cysteine conjugate were excreted in the urine of males and females, whereas the sulfate and/or glucuronide conjugate of the OH-phenyl metabolite was seen only in the urine of male rats. Based on rapid elimination, diclosulam has little potential to accumulate upon repeated administration (Stewart et al., 1996).

87.3.3  Toxicity to Humans No studies are available on intentional human exposure. However, risk assessments using conservative assumptions indicate high margins of safety with diclosulam. Residue studies indicated no detectable residues at a limit of detection of 0.003 ppm, and no likelihood for accumulation. A tolerance level of 0.02 ppm, based on a limit of quantitation of 0.01 ppm, and a reference dose of 0.05 mg/kg/day based on the lowest NOEL (5 mg/kg/day from the chronic rat study) have been proposed (U.S. EPA, 1998). Calculation of a maximum potential average daily dose assuming 100% of proposed crops with residues equal to the tolerance level indicates theoretical exposure to only 0.1% of the RfD in the population with the highest potential expos­ ure (nonnursing infants under 1 year old). The MOE for occupational exposure to diclosulam, calculated using exposure estimates from the U.S. EPA Pesticide Handlers Exposure Database (PHED), is estimated to be greater than 1000 based on the NOEL from the chronic dog study and assuming 100% absorption.

87.4  Florasulam 87.4.1  Identity, Properties, and Uses 87.4.1.1  Chemical Name The IUPAC name for florasulam is 2,6,8-trifluoro-5-me thoxy[1,2,4]triazolo[1,5-c]pyrimidine-2-sulfonanilide; the CAS name is N-(2,6-difluorophenyl)-8-fluoro-5-methoxy [1,2,4]triazolo[1,5-c]pyrimidine-2-sulfonamide.

Chapter | 87  Toxicology of Triazolopyrimidine Herbicides

87.4.1.2  Structure See Figure 87.1 and Table 87.1.

87.4.1.3  Synonyms Florasulam is also known as XR-570, XDE-570, or DE570 and is sold either alone or in combination under registered trade names including Primus, Derby, Kantor, and Mustang herbicides. The CAS registry number is 145701-23-1.

87.4.1.4  Physical and Chemical Properties Florasulam is a light-colored solid with an empirical formula of C12H8F3N5O3S and a molecular weight of 359.3. It has a low vapor pressure (7.5  108 mmHg at 25°C) and decomposes at 193.5–230.5°C. The solubility increases with increasing pH, ranging from 84 mg/l at pH 5 to 9400 mg/l at pH 9. Florasulam is highly soluble in acetone (123 g/l) and acetonitrile (72 g/l), but substantially less soluble in octanol (0.18 g/l) and xylene (0.23 g/l). It has a pKa of 4.54 and log Kow values ranging from 1.00 at pH 4 to 2.06 at pH 10.

87.4.1.5  Uses Florasulam is a highly effective postemergence broadleaf herbicide for use in cereals, grassland, and turf. Maximum label use rate for the various crops range from 5 to 10 g per hectare.

87.4.2  Toxicity to Laboratory Animals 87.4.2.1  Acute Exposure Florasulam was essentially nonhazardous by the oral, dermal, and inhalation routes, was nonirritating to skin and eyes, and did not induce delayed contact hypersensitivity in either a modified Buehler test or a Magnusson and Kligman maximization study. The oral LD50 was greater than 6000 mg/kg in the rat and 5000 mg/kg/day in the mouse, the dermal LD50 in the rabbit was greater than 2000 mg/kg, and the 4-h inhal­ ation LC50 in the rat exceeded 5 mg/l (Brooks, 1997; Clements and Cieszlak, 1995; Gilbert, 1995a,b,c,d; Gilbert and Yano, 1995b; Johnson, 1996).

87.4.2.2  Repeated Exposure In dietary studies of 2- to 13-week duration, the kidney was identified as a target organ in rats, mice, and dogs, whereas the liver was also a target organ in dogs. In F344 rats, subacute exposure to dosages of 500 mg/ kg/day and above was associated with karyomegaly and anisokaryocytosis in proximal tubular epithelial cells in males and females, and tubular degeneration with regeneration in females. Individual proximal tubular cell necrosis

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was also seen in both sexes at 1000 mg/kg/day. The NOEL was 100 mg/kg/day. Subchronic studies were conducted in F344 and Sprague-Dawley rats at dosages up to 1000 mg/ kg/day. Dosages of 500 or 1000 mg/kg/day produced necrosis with regeneration in descending proximal tubules and a marginally increased incidence of degeneration with regeneration of renal tubules in females. Papillary mineralization (tubular debris) and papillary necrosis were reported at the highest dosages (800 mg/kg/day). Other highdosage effects included acidic urine (males only), increased kidney weight, perineal soiling, reduced body weight gain and feed consumption (due, at least in part, to reduced palatability of the diet), and reduced red blood cell indices. With the exception of mineralized debris in renal papillae and degeneration and regeneration of cortical tubules, all effects partially or completely resolved by the end of a 4-week recovery period. The subchronic NOEL in rats was 100 mg/kg/day (Liberacki et al., 1996; Redmond and Johnson, 1996a; Szabo and Davis, 1993b). In B6C3F1 mice, subacute exposure to dosages up to 1000 mg/kg/day was without effect (Szabo and Davis, 1992). The only response to subchronic exposure to dosages up to 1000 mg/kg/day was hypertrophy of renal collecting duct epithelial cells at 500 mg/kg/day and above. The subchronic NOEL in mice was 100 mg/kg/day (Redmond and Johnson, 1996b). In Beagle dogs, subacute exposure to a nominal dosage of 450 mg/kg/day was associated with reduced body weight gain and reduced feed consumption (due, at least in part, to reduced palatability of the diet). Hepatic changes characterized by increased liver weight and bile duct hyperplasia in males and females, and bile stasis and hep­ atocellular necrosis in males were observed in this group. At 150 mg/kg/day, the effects were limited to increased liver weight and bile duct hyperplasia. Serum alkaline phosphatase (AP) activity, probably of hepatic origin, was elevated at all dosages, including the low dosage of 50 mg/ kg/day. The increase in serum AP at the low dosage was without histopathological correlate. Therefore, 50 mg/kg/ day was considered a subacute NOAEL. Liver effects were not exacerbated by an extended treatment period, but renal hypertrophy (not seen after 4 weeks of exposure) similar to that reported in rats and mice was evident in dogs after subchronic exposure to 50 mg/kg/day (Sullivan and CroninSingleton, 1995; Sullivan and Singleton, 1995). Repeated dermal exposure to dosages up to 1000 mg/ kg/day for 4 weeks produced only transient dermal irritation in rats during the last week of treatment, with no systemic effects (Scortichini and Kociba, 1997).

87.4.2.3  Chronic Toxicity and Carcinogenicity In chronic dietary studies (1- to 2-year duration), hyper­trophy of a population of renal collecting duct cells identified as -intercalated cells remained the most sensitive ­morphological

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effect in all species. Hypertrophy was present at 50 mg/kg/day in dogs, 125 and 250 mg/kg/day and above in male and female rats, respectively, and 500 mg/kg/day and above in mice. In F344 rats, 2-year dietary exposure to dosages between 10 and 500 mg/kg/day identified the kidney as the only target organ. At the high-dose level (250 mg/kg/day in females and 500 mg/kg/day in males), very slight to slight hyper­trophy of renal collecting duct epithelial cells was evident after 1 year. After 2 years, this change had progressed to a moderate degree in some males. Other effects at this dose level included reduced body weight gain, reduced urinary pH, perineal soiling, reduced red blood cell indices, and renal changes similar to effects seen following subchronic expos­ ure to high-dose levels. At the next dose level (125 mg/kg/ day in females and 250 mg/kg/day in males), very slight to slight hypertrophy of renal collecting duct epithelial cells was evident in males after 1 year, and in males and females after 2 years. Reduced body weight gain, renal papillary mineralization (males), decreases in spontaneous chronic renal disease, reduced urinary pH, and perineal soiling were also seen in these animals. No effects occurred at 10 mg/kg/ day and there was no treatment-related effect on tumor formation in this study (Johnson et al., 1997). In B6C3F1 mice, 2-year dietary exposure to dosages up to 1000 mg/kg/day identified the kidney as the only target organ with hypertrophy of intercalated cells, decreased renal epithelial cell cytoplasmic lipid-like microvacuoles, and a decreased incidence of spontaneous chronic renal disease at 1000 mg/kg/day. Reduced body weights accompanied by minor changes in serum cholesterol and triglycerides were also present at 1000 mg/kg/day. At 500 mg/kg/day, very slight hypertrophy of renal collecting duct epithelial cells occurred in most males and females along with decreases in cytoplasmic lipid-like microvacuoles and spontaneous chronic disease (females only) of renal tubules. No effects occurred at 50 mg/kg/day, and there was no treatment-related tumorigenic or carcinogenic response at any dose level (Quast et al., 1997). In Beagle dogs, 1-year dietary exposure to florasulam revealed kidneys, liver, and adrenal glands as target organs. In the subchronic study, renal hypertrophy and modest elevations of serum alkaline phosphatase and liver weight were the only treatment-related effects seen at 100 mg/kg/day. However, treatment beyond 13 weeks resulted in reduced food consumption and body weight gain in some animals, and significant elevations in serum enzyme activities associated with liver toxicity. The original high dosage of 100 mg/ kg/day was therefore reduced to 50 mg/kg/day on Day 105. Thereafter, food consumption and body weight gain improved and red blood cell indices in females and serum transaminases in both sexes returned to normal. Serum alkaline phosphatase remained elevated to the end of the study. Slight hypertrophy of renal collecting duct epithelial cells and slight vacuolization of the zona reticularis and zona

Hayes’ Handbook of Pesticide Toxicology

fasciculata of adrenal glands were detected histologically in this high-dosage group. The fatty change in the adrenals of dogs represented a slight exacerbation of a spontaneous lesion, not associated with inflammation, necrosis, or clinical chemistry changes, and was considered of uncertain toxicological importance. No histopathological lesions were evident in the liver. The NOEL was 5 mg/kg/day (Stebbins and Haut, 1997). Histological and ultrastructural evaluation of the affected renal collecting duct cells characterized the hypertrophy as a mitochondrial proliferation of -intercalated cells, which functionally are involved in acid-base regulation and contain high levels of H-ATPase and HK-ATPase in the apical membrane (Brown et al., 1988; Garg, 1991; Hamm and Hering-Smith, 1993; Madsen and Tisher, 1986; Stokes, 1993; Verlander et al., 1991). Hypertrophy of intercalated cells has been reported as a physiological response to several factors affecting acid-base homeostasis, including respiratory acid­osis, metabolic acidosis, hypokalemia, and altered serum ad­renal mineralocorticoid levels (Ahn et al., 1996a,b; DeFronzo, 1980; Eiam-ong et al., 1994; Hansen et al., 1980; Madsen et al., 1991; Tsuruoka and Schwartz, 1996a,b; Verlander et al., 1994; Weiner and Wingo, 1997; Wingo and Cain, 1993). However, none of these factors was found to be adversely affected by florasulam (Weiner, 1997). The lack of any adverse sequelae associated with this change suggests that it is an adaptive rather than an adverse response to florasulam.

87.4.2.4  Mutagenicity In a battery of tests, florasulam showed no evidence of genotoxic potential. These tests included an in vitro bacterial reverse mutation assay (Ames test), an in vitro cytogenetic assay in Chinese hamster ovary cells (CHO/ HGPRT assay), an in vitro chromosomal aberration assay in rat lymphocytes, and an in vivo cytogenetic assay in mouse bone marrow cells (Lawlor, 1995; Lick et al., 1995; Linscombe et al., 1995a,b).

87.4.2.5  Neurotoxicity Acute gavage and chronic (1-year dietary) neurotoxicity studies in Fischer 344 rats revealed only nonspecific findings. In both acute and chronic neurotoxicity studies with florasulam, perineal urine staining at the highest dosages was the only treatment-related effect. No other effects were seen following an extensive battery of neurologic tests and neurohistopathological examinations (Mattsson and McGuirk, 1997; Shankar and Johnson, 1996).

87.4.2.6  Reproductive Toxicity In developmental toxicity studies, there were no adverse effects on intrauterine development or prenatal survival

Chapter | 87  Toxicology of Triazolopyrimidine Herbicides

in rats or rabbits administered gavage dosages as high as 600–750 mg/kg/day. Maternal effects on survival, feed consumption, and/or weight gains occurred at these high dosages. In SD rats, the embryo-fetal NOEL was 750 mg/ kg/day, whereas the maternal NOEL was 250 mg/kg/day. In New Zealand White rabbits, the NOEL for both maternal and embryo-fetal effects was 500 mg/kg/day. In a twogeneration dietary reproduction study in SD rats at dosages of 10–500 mg/kg/day, parental effects (weight gain, feed consumption, renal changes) were seen only at the highest dosage, with no effects on any reproductive parameter. Transient decreases in neonatal body weights, secondary to decreases in maternal feed consumption, were seen at 500 mg/kg/day. The parental NOEL was 100 mg/kg/day, whereas the NOEL for reproductive effects was 500 mg/kg/ day (Liberacki and Carney, 1997; Liberacki et al., 1997; Zablotny and Carney, 1997).

87.4.2.7  Absorption, Distribution, Metabolism, and Excretion In metabolism studies in F344 rats, single oral doses of 10–500 mg/kg 14C-florasulam were readily and extensively absorbed (90% of a 10-mg/kg dose within 24 h) and rapidly eliminated (plasma t1/2  8–10 h) primarily in the urine (85% of administered dose). The feces contained small amounts of the administered radioactivity (5–17%) depending on the dose. More than 75% of the 14C activity in urine was found to be unchanged florasulam. Two minor metabolites, identified as a free and a conjugated (sulfated) hydro­xyphenyl derivative of florasulam, were found. Feces contained unchanged florasulam and the free hydroxyphenyl-derivative. There was no evidence of hydrolysis of the sulfonamide bridge based on the metabolites found in the urine and feces. The rapid elimination of florasulam from tissues indicated no potential to accumulate upon repeated administration (Dryzga et al., 1996; Hansen, 1997). Absorption following in vivo dermal expos­ ure of rats to a concentrated suspension formulation containing 14C-florasulam was minimal (mean 0.5% over 72 h). Results obtained from in vitro studies were similar to those obtained from the in vivo study (Bounds, 1997; Perkins and Billington, 1998).

87.4.3  Toxicity to Humans No studies are available on intentional human exposure. Risk assessment calculations for the general population and for pesticide handlers indicate a low-risk estimate. Residue studies have indicated no detectable levels in cereal gains at the limit of quantitation. Maximum residue levels (MRLs) of 0.01 ppm in grains, and 0.05 ppm whole plants and straw based on the limit of detection, and a reference dose of 0.05 mg/kg/day on the basis of a chronic NOEL of 5 mg/kg/day in dogs have been proposed. The

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theoretical dietary intake of florasulam from all routes has been estimated to account for 0.1% of the ADI even in children 1–2 years of age, the most highly exposed population subgroup. Margins of Exposure (MOE) for pesticide handlers were determined to be significantly greater than 100 and thus florasulam does not pose a risk for occupational exposure (U.S. EPA, 2007a).

87.5  Flumetsulam 87.5.1  Identity, Properties, and Uses 87.5.1.1  Chemical Name The IUPAC name for flumetsulam is 2,6-difluoro-5-meth yl[1,2,4]triazolo[1,5-a]pyrimidine-2-sulfonanilide; the CAS name is N-(2,6-difluorophenyl)-5-methyl[1,2,4]triazolo[1, 5-a]pyrimidine-2-sulfonamide.

87.5.1.2  Structure See Figure 87.1 and Table 87.1.

87.5.1.3  Synonyms Flumetsulam (also known as XRD-498) is the generic name for this material which is sold globally under registered trade names including Broadstrike, Python, Preside, and Scorpion herbicides. The CAS registry number is 98967-40-9.

87.5.1.4  Physical and Chemical Properties The empirical formula of flumetsulam is C12H9F2N5O2S, with a molecular weight of 325.3. Flumetsulam is a lightcolored powder at room temperature, with a melting point of 252.9°C, a vapor pressure of 2.8  1015 mmHg at 25°C, a Kow of 1.62 at pH 3.44, and a pKa of 4.60. Flumetsulam is soluble in water at 5.65 g/l at pH 7 and 25°C, but solubility decreases with decreasing pH, and it is less soluble in organic solvents.

87.5.1.5  Uses Flumetsulam is a broad-spectrum, season-long herbicide used in the control of broadleaf weeds in soybeans, corn, and other major crops. Flumetsulam is applied as a soilincorporated preplanting, preemergence, or postemergence herbicide depending on the formulation, at a maximum use rate of 80 g per hectare.

87.5.2  Toxicity to Laboratory Animals 87.5.2.1  Acute Exposure Flumetsulam is essentially nontoxic following acute exposure, with an acute oral LD50 greater than 5000 mg/kg, a dermal LD50 greater than 2000 mg/kg, an acute 4-h

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inhalation LC50 above the highest attainable aerosol concentration of 1.2 mg/l of air. It is a slight, transient eye irritant and produced no signs of dermal irritation in rabbits following acute exposure, nor was there any evidence of dermal sensitization in guinea pigs (Mizell et al. 1988a,b, c,d,e; U.S. EPA, 1993).

levels, with no histopathologic correlates, were reported in both males and females. A dosage of 500 mg/kg/day was considered the NOAEL in females (Cosse et al., 1989). Repeated dermal exposure to dosages of 100 mg/kg/day for 21 days produced very slight epidermal hyperplasia, but no indications of any systemic effects (Stebbins et al., 1990).

87.5.2.2  Repeated Exposure

87.5.2.3  Chronic Toxicity and Carcinogenicity

Subchronic toxicity studies in rats, mice, and dogs indicated a low degree of toxicity following repeated oral exposure. In rats, exposure to dietary concentrations of up to 5% (approximately 6000 mg/kg/day) for 2–4 weeks identified the kidney as the primary target organ. Effects in the kidney consisted of focal necrosis and inflammation of the papilla(e), and tubular epithelial cell degeneration and regeneration, with secondary effects on urinalysis parameters at the highest dosage. The only effect reported at 1000 mg/kg/day was cecal enlargement in males. However, the ceca were histologically normal, and 1000 mg/kg/day was considered the NOAEL following 2–4 weeks of exposure. Rats fed diets containing flumetsulam at dosages of 250–2500 mg/kg/day for 13 weeks exhibited dose-dependent changes similar to those seen after 4 weeks of exposure. The NOAEL in rats following subchronic exposure was 25 mg/kg/day (Yano et al., 1987, 1988; Zempel et al., 1988). In B6C3F1 mice fed flumetsulam for 2 weeks, decreased kidney weights were reported in males at dietary concentrations of 1.5 and 3.0% and in females given 3.0%, which corresponded to dosages 3500 mg/kg/day. The NOEL in mice was 0.5% (approximately 1150–1365 mg/kg/day). B6C3F1 mice given dosages of 100–5000 mg/kg/day for 13 weeks displayed only a minimal increase in centrolobular-to-midzonal hepatocellular eosinophilia at the highest dose level and decreased vacuolation of renal proximal tubular epithelium, which is of doubtful toxicologic significance. The NOEL for mice was 1000 mg/kg/day, and 5000 mg/ kg/day was considered a NOAEL (Bond et al., 1987; Stott et al., 1986). In both rats and mice, the increases in the size and weight of the cecum, observed only at high dosages and unassociated with any histologic changes, were considered adaptive in nature, most likely secondary to the effects of flumetsulam on the microenvironment within the cecum. Beagle dogs were fed flumetsulam at nominal dosages of 100–1000 mg/kg/day (males) or 1500 or 2500 mg/kg/day (females) for 2 weeks. In females, degeneration and regeneration of the renal tubular epithelial cells and lymphocytic infiltration of hepatic sinusoids were reported. The NOEL in dogs was approximately 800 mg/kg/day (nominally 1000 mg/kg/day). In dogs given a dosage of 1000 mg/kg/day for 13 weeks, degenerative microscopic changes in the renal papilla, slight biliary stasis, and hep­atocellular necrosis were observed. At 500 mg/kg/day, slight renal papillary degeneration was noted microscopically in males, and increases in serum AP and globulin levels and decreased serum albumin

Flumetsulam was fed to F344 rats and B6C3F1 mice for 2 years at dosages of 100–1000 mg/kg/day. No treatmentrelated adverse effects were noted in mice (Bond et al., 1991). In rats, atrophy of the renal papilla(e) with secondary hyperplasia and/or mineralization of the pelvic epithelium were noted in males given 1000 mg/kg/day, but not in females, and the NOEL was 500 mg/kg/day (Stott et al., 1991). As was noted following subchronic exposure, cecal enlargement with no accompanying histopathologic changes was observed in both rats and mice following chronic exposure. There was no evidence of a tumorigenic or carcinogenic response in either rats or mice at dosages up to 1000 mg/kg/day. The dog appeared to be the most sensitive species to long-term exposure to flumetsulam. Administration of dosages of 500 mg/kg/day in the diet for 1 year produced inflammatory and atrophic changes in the kidney, accompanied by calculi in females. At 100 mg/kg/day, only increased alkaline phosphatase activity and decreased serum albumin were reported, with no histologic changes in any organs. The chronic NOEL in dogs from this study was 20 mg/kg/day, whereas the NOAEL was 100 mg/kg/day (Yano et al., 1991).

87.5.2.4  Mutagenicity In a battery of tests, flumetsulam showed no evidence of genotoxic potential. Flumetsulam was negative in an in vitro bacterial reverse mutation assay (Ames test), an in vitro cytogenetic assay in Chinese hamster ovary cells (CHO/HGPRT assay), an in vitro rat hepatocyte unscheduled DNA synthesis (UDS) assay, and an in vivo cytogenetic assay in mouse bone marrow cells (U.S. EPA, 1993).

87.5.2.5  Reproductive Toxicity Flumetsulam did not affect development or reproduction in either rats or rabbits. No evidence of maternal toxicity, embryo-fetotoxicity, or teratogenicity was observed in rats following exposure of pregnant females to 1000 mg/kg/day in the diet, though the weights of the ceca were increased, consistent with effects noted in previous dietary studies. No parental toxicity or alterations in reproductive performance occurred in rats given up to 1000 mg/kg/day over two generations. Gavage administration of flumetsulam to pregnant rabbits at dosages of 500–700 mg/kg/day produced dose-related episodes of anorexia, with sequelae

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secondary to the altered nutritional status (deteriorated clinical condition, mortality, stomach erosions, etc.), but no embryo-fetotoxicity or teratogenicity accompanied these maternal effects. The maternal NOEL from this study was 100 mg/kg/day, whereas the NOEL for embryo-fetal development was 700 mg/kg/day (Hanley, 1989; Zempel et al., 1990; Zielke et al., 1988).

sulfonanilide; the CAS name is N-(2,6-dichloro-3-methylphenyl)-5,7-dimethoxy[1,2,4]triazolo[1,5-a]pyrimidine-2sulfonamide.

87.5.2.6  Absorption, Distribution, Metabolism, and Excretion

87.6.1.3  Synonyms

Flumetsulam was rapidly, though incompletely, absorbed in mice and rats, with absorption half-lives of less than 1 h following oral administration of doses of either 5 or 1000 mg/kg. Excretion was also rapid, with a urinary halflife of approximately 5–7 h. Following oral administration of 14C-flumetsulam, approximately 50–75% of the administered radiolabel was excreted in the urine primarily as unchanged parent material, though two minor (20% of urinary radiolabel) metabolites, believed to be conjugates of parent flumetsulam, were found in the urine of mice. Approximately 20–35% of the dose was found in the feces, which represented apparently unabsorbed flumetsulam (based on almost total elimination in the urine of an intravenous dose to rats) and tissue levels of 14C accounted for less than 1.5% of the administered dose. There were no differences in absorption, distribution or elimination based on sex, though slight differences were seen with increasing dose (Pottenger et al., 1991; Timchalk et al., 1988).

87.5.3  Toxicity to Humans No studies are available on intentional human exposure. Risk assessment indicates a low potential risk from normal use of flumetsulam. Residue tolerances of 0.05 ppm have been set by the U.S. EPA for soybeans and for corn grain, fodder, and forage. A reference dose of 1 mg/kg/day was established on the basis of a NOAEL of 100 mg/kg/ day from a 1-year study in dogs. Dietary risk evaluation assuming 100% of crops are treated and residues are at the established tolerance levels indicates only 0.2% of the RfD is used even by the highest exposed subgroup, children 3–5 years (U.S. EPA, 2008). Based on the use rates and the NOEL from the chronic dog study, the MOE for worker exposure is greater than 100 and as such is deemed not to represent a risk to humans.

87.6  Metosulam 87.6.1  Identity, Properties, and Uses 87.6.1.1  Chemical Name The IUPAC name for metosulam is 2,6-dichloro-5,7dimethoxy-3-methyl[1,2,4]triazolo[1,5-a] pyrimidine-2-

87.6.1.2  Structure See Figure 87.1 and Table 87.1.

Metosulam is also known as methoxsulam, XRD-511, XDE-511, and DE-511, and is sold either alone or in combination, under a variety of registered trade names including Tacco, Sansac, Eclipse, Atol, Kompal, and Sinal herbicides. The CAS number is 139528-85-1.

87.6.1.4  Physical and Chemical Properties Metosulam is a cream- to tan-colored powder with a low vapor pressure (7.5  1015 mmHg at 25°C). The empir­ ical formula is C14H13Cl2N5O4S, and the molecular weight is 418.3. The solubility of metosulam in water at 20°C and pH 7 is 700 mg/l. Given a pKa of 4.8, the solubility is pHdependent, with values of 100 mg/l at pH 5 and 5600 mg/l at pH 9 (at 20°C), and the log Kow is 2.12 at pH 5.

87.6.1.5  Uses Metosulam is a broad-spectrum, postemergence broadleaf herbicide intended for use in cereals, maize, pasture, alfalfa, and rice. Maximum label use rates for the various crops range from 5 to 30 g per hectare.

87.6.2  Toxicity to Laboratory Animals 87.6.2.1  Acute Exposure Metosulam has very low acute toxicity. The oral LD50 in Fischer 344 rats and CD-1 mice is greater than 5000 mg/kg. No toxicity, including histopathological changes of eyes and kidneys, was evident in beagle dogs given one to five daily doses of 2000 mg/kg, by gelatin capsule. The dermal LD50 in the rabbit was greater than 2000 mg/kg. The 4-h inhal­ ation LC50 in the rat was greater than the highest attainable concentration of 1.9 mg/l of air. Metosulam, when applied to the intact skin of rabbits, produced no signs of irritation. Following instillation into rabbit eyes, slight conjunctival redness developed within 1 h of treatment, but all treated eyes were normal within 1 day of treatment. There was no indication of contact sensitization in guinea pigs exposed to metosulam using either a Magnusson and Kligman maximization test or a modified Buehler topical patch method (Lockwood, 1989a,b,c; Lockwood and Szabo, 1989a,b; UK Ministry of Agriculture, Fisheries, and Food, 1996).

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87.6.2.2  Repeated Exposure

87.6.2.3  Chronic Toxicity and Carcinogenicity

Repeated exposure toxicity studies were conducted with metosulam in rats, mice, dogs, rabbits, and nonhuman primates (cynomolgus monkeys). In rats, dietary administration of dosages of 500–5000 mg/kg/day for 2 weeks to Sprague-Dawley rats resulted in lower body weights associated with unpalatability and the NOEL was 100 mg/kg/ day. No significant effects were reported in Long Evans rats administered dosages of up to 2000 mg/kg/day for 2 weeks. Following subchronic exposure, the kidney was identified as the major target organ. In a 13-week dietary study, the primary toxicological effects were renal alterations characterized as hypertrophy and nuclear pleomorphism of cells lining the proximal convoluted tubules at 100 mg/kg/day and above. After 4 weeks on control diet, hypertrophy of renal tubular cells had resolved, and nuclear pleomorphism was markedly decreased. The NOEL for subchronic dietary administration of metosulam in rats was 10 mg/kg/day. CD-1 mice administered metosulam in the diet at 100– 5000 mg/kg/day for 2 weeks exhibited centrilobular hepatocellular necrosis and decreased vacuolation in the liver only at 2000 mg/kg/day and above. The NOEL was 1000 mg/kg/ day. The only effect observed in a 13-week dietary study at dosages up to 2000 mg/kg/day, was mild hepatocellular hypertrophy and the NOEL was 250 mg/kg/day. The kidney was identified as the most sensitive target organ in the dog (similar to the rat). In addition, ocular toxicity in the form of retinal damage unique to this species was also observed. Dietary administration of metosulam to Beagle dogs at dosages of 100–1000 mg/kg/day for 14 days produced dose-related decreases in feed consumption and body weights; retinal degeneration, necrosis, and detachment; and degeneration or focal necrosis of distal renal collecting tubules and collecting ducts. Metosulam was fed to dogs at dosages of 5, 25, and 50 mg/kg/day for 13 weeks. Clinical signs of blindness occurred as early as 6 weeks in all dogs administered 50 mg/kg/day. Microscopic examination of the eyes from these dogs showed retinal degeneration with detachment. Choroidal structures (tapetum lucidum, pigmented epithelium, and choroidal blood vessels) and other ocular structures were normal. Ocular tissues from dogs administered 5 mg/kg/ day were normal. In the kidneys, very slight to moderate degeneration of the distal convoluted tubules and collecting ducts of dogs administered 25 mg/kg/day and above was reported, and the NOEL was 5 mg/kg/day. Male and female cynomolgus monkeys exposed to oral dosages of 0 or 100 mg/kg/day for 6 weeks showed no renal or ocular toxicity after detailed examination which included an extensive histopathologic evaluation. Repeated dermal exposure of New Zealand White rabbits to dosages of up to 1000 mg/kg/day for 21 days produced no signs of dermal irritation or systemic effects (UK Ministry of Agriculture, Fisheries, and Food, 1996).

Following chronic (2-y) exposure in Sprague-Dawley rats at dosages of 5–100 mg/kg/day, the primary effects were confined to the kidneys, consistent with the findings following subchronic exposure, and the effects were more severe in male than in female rats. At 100 mg/kg/day, nuclear pleomorphism and hyperplasia of cells of the proximal tubules as well as basophilic adenomas and adenocarcinomas of the renal cortex were observed. At 30 mg/ kg/day, nuclear pleomorphism of proximal tubular cells was present and only a single renal cortical adenocarcinoma, which was within the historical control incidence for this tumor (Charles River Breeding Laboratories, 1987), was observed in this group. Short-term exposure studies demonstrated the presence of mitotic figures and nuclear pleomorphism in the renal cortex of male rats following as little as 1 week of dietary exposure to 100 mg metosulam/ kg/day. Increased mitotic activity measured by BrdU incorporation correlated with the renal tubular epithelial changes noted histologically (UK Ministry of Agriculture, Fisheries, and Food, 1996). This suggested a nongenotoxic mechanism of repeated injury as described by Dietrich and Swenberg (1991) as the probable origin of the renal tumors in the chronic study with metosulam. In CD-1 mice fed dose levels of metosulam of up to 1000 mg/kg/day for 18 months, there was no evidence of any increase in tumor incidence and no effects were noted in any other parameter. Metosulam administered to Beagle dogs at dosages of 3–37.5 mg/kg/day for 12 months produced effects in the eyes and kidneys consistent with the findings of the subchronic study. At 37.5 mg/kg/day, variable retinal degeneration with detachment, beginning with diminished or absent pupillary light reflex, increased tapetal reflectivity, and progressive retinal deterioration, were observed. Degenerative lesions of the distal convoluted tubules and collecting ducts were also observed at 37.5 mg/kg/day. No effects were observed at lower levels, and the NOEL was 10 mg/kg/day (UK Ministry of Agriculture, Fisheries, and Food, 1996). The sensitivity of the dog eye to metosulam appears to be unique to this species. The pathology involved the loss of the photoreceptor layer and its nuclei, together with a collapse of the outer and inner plexiform layers and inner nuclear layer. No retinopathy was associated with the pigmented epithelial layer or in the tapetal cells. It is important to note that retinal pathologies were not detected with metosulam in any other species. Dosages of 300 mg/kg/day for 12 days in rabbits, and up to 1000 and 2000 mg/kg/ day for 13 weeks in rats and mice, respectively, were not associated with any retinal changes. Exposure of mice to 1000 mg/kg/day for 18 months or rats to 100 mg/kg/day for 2 years likewise induced no retinal pathology. Significantly, dosages of 100 mg/kg/day for 6 weeks in the nonhuman primate, the species most closely resembling human ocular

Chapter | 87  Toxicology of Triazolopyrimidine Herbicides

anatomy and physiology, also produced no evidence of retinal toxicity. Pharmacokinetic studies using radiolabeled metosulam indicated metosulam localized over the outer layer of the retina in the dog, but no selective localization was detected in the rat or mouse (see below) (UK Ministry of Agriculture, Fisheries, and Food, 1996).

87.6.2.4  Mutagenicity In a battery of tests, metosulam showed no evidence of genotoxic potential. These tests included an in vitro bacter­ ial reverse mutation assay (Ames test), an in vitro mammalian forward mutation test in Chinese hamster ovary cells (CHO/HGPRT assay), a mammalian cytogenetics test in rat lymphocytes, an in vitro unscheduled DNA synthesis (UDS) assay, and an in vivo mouse bone marrow micronucleus test (UK Ministry of Agriculture, Fisheries, and Food, 1996).

87.6.2.5  Reproductive Toxicity Metosulam did not produce any adverse reproductive or developmental effects when tested in rats and rabbits. There were no effects on maternal or developmental parameters in a conventional teratogenicity study in the SD rat at dietary levels up to 1000 mg/kg/day. In New Zealand White rabbits, maternal effects were noted at oral gavage dosages of 100 or 300 mg/kg/day and the maternal NOEL in rabbits was 30 mg/kg/day, but there was no indication of developmental effects at 300 mg/kg/day. In a two­generation reproduction study in Sprague-Dawley rats at dosages of 5–100 mg/kg/day, renal toxicity was observed among the parental rats at 100 mg/kg/day, consistent with the effects noted following chronic exposure, but reproductive performance was unaffected. The NOEL for parental toxicity from this study was 30 mg/kg/day, whereas the NOEL for reproductive effects was 100 mg/kg/day (UK Ministry of Agriculture, Fisheries, and Food, 1996).

87.6.2.6  Absorption, Distribution, Metabolism, and Excretion The metabolic fate of 14C-metosulam in rats, mice, and dogs following single or multiple oral administrations was evaluated. 14C-Metosulam was absorbed rapidly (t1/2  1 h) in all three species, though the extent of absorption was significantly higher in the rat (70%) than in the dog and mouse (20%). The rate of 14C elimination in mice and rats was comparable (t1/2  54–60 h), whereas the elimin­ ation rate in dogs was slightly slower (t1/2  73 h). In all three species, 14C-metosulam and metabolites were excreted in the urine. HPLC analysis of urine samples revealed extensive metabolism in both mice and rats, but much less pronounced metabolism in dogs. Analysis of 14C activity of the dog eyes indicated that this organ, a target for toxicity in the dog, exhibited an affinity for the radiotracer not seen in other species. Histoautoradiographic sections

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of dog eyes revealed radioactivity localized regionally over the outer layer of the retina, whereas analysis of tissues from rats and mice for 14C activity and histoautoradiography indicated a lack of selective affinity for any ocular tissues (Timchalk et al., 1996). The major metabolites were an oxidation product of the 3-methyl moiety of the phenyl ring and a demethylation of the 3-methoxy moiety of the pyrimidine ring. In studies conducted in male rats with 14Cmetosulam labeled in either the phenyl or pyrimidine ring, no evidence of cleavage of the sulfonamide bridge was seen (UK Ministry of Agriculture, Fisheries, and Food, 1996). In vitro dermal penetration studies using rat (SpragueDawley) and fresh human skin demonstrate that less than 1% of the applied metosulam actually penetrated the skin (UK Ministry of Agriculture, Fisheries, and Food, 1996).

87.6.3  Toxicity to Humans No studies are available on intentional human exposure. Risk assessment calculations for the general population and for pesticide handlers indicate acceptable risk estimates. Residue studies have indicated no detectable levels in cereal grains at the limit of quantitation. Maximum residue levels of 0.1 ppm in grains based on the limit of quantitation, and an acceptable daily intake (ADI) of 0.01 mg/kg/day on the basis of the chronic NOEL of 5 mg/ kg/day in rats and a conservative safety factor of 500 have been proposed. Using these values, the maximum theoretical dietary intake (MTDI) of metosulam from all routes of exposure has been estimated to account for 5% of the ADI. An acceptable operator exposure level (AOEL) of 0.04 mg/kg/day has been calculated using a NOEL of 10 mg/kg/day and a safety factor of 250 (UK Ministry of Agriculture, Fisheries, and Food, 1996).

87.7  Penoxsulam 87.7.1  Identity, Properties, and Uses 87.7.1.1  Chemical Name The IUPAC name for penoxsulam is 3-(2,2-difluoroethoxy)-N-(5,8-dimethoxy[1,2,4]triazolo[1,5-c]pyrimidin2-yl)-,,-trifluorotoluene-2-sulfonamide; the CAS name is 2-(2,2-difluoroethoxy)-N-(5,8-dimethoxy[1,2,4]triazolo[1,5-c] pyrimidin-2-yl)-6-(trifluoromethyl)benzenesulfonamide.

87.7.1.2  Structure See Figure 87.1 and Table 87.1.

87.7.1.3  Synonyms Penoxsulam is also known as XR-638, XDE-638, and DE-638 and is sold either alone or in combination, under

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a variety of registered trade names including: Citadel, Granite, Ricer, and Rainbow herbicides. The CAS number is 219714-96-2.

87.7.1.4  Physical and Chemical Properties Penoxsulam is an off-white-colored solid. It has a vapor pressure of 1.87 3 10216 mmHG at 20°C. The empirical formula is C16H14F5N5O5S and the molecular weight is 483.37. The solubility of penoxsulam in water at 19°C and pH 7 is 0.408 g/l. At pH 5, solubility is 0.00566 and 1.46 at pH 9.0. The log Kow in unbuffered water at 19°C is 0.354.

87.7.1.5  Uses Penoxsulam is a broad-spectrum systemic broadleaf herbi­ cide with pre- and postemergent uses. It was originally developed as a rice herbicide and is also registered for use on turf, trees, and vines and control of aquatic vegetation. Due to low use rates (17.535 g per hectare) and favorable environmental and human health profiles, penoxsulam has been designated as a reduced-risk pesticide for these uses by the U.S. EPA.

87.7.2  Toxicity to Laboratory Animals 87.7.2.1  Acute Exposure Penoxsulam has very low acute toxicity by the oral, dermal, and inhalation routes. It caused mild eye irritation and only very slight dermal irritation and was not a skin sensitizer (Magnusson and Kligman maximization method). The oral and dermal LD50 values were each greater than 5000 mg/kg in rats while the 4-h inhalation LC50 was 3.50 mg/l, the highest technically attainable aerosol concentration (Bonnette, 2000a,b,c,d,e; Hoffman, 1999).

87.7.2.2  Repeated Exposure The liver and/or kidneys were identified as target organs in rats, mice and dogs following dietary exposure for 4- and 13-week treatment intervals. In a 4-week study in Fischer 344 (F344) rats, urine soiling was noted in females at 500 and 1000 mg/kg/day. Body weight gains were reduced in both sexes at doses 500 mg/ kg/day. Liver (males and females) and kidney (females) weights were also increased at 500 mg/kg/day and multi­ focal hyperplasia and inflammation of the renal pelvic epithelium was observed in females at the same dose levels. Based on these findings, the NOAEL was established at 100 mg/kg/day. Subchronic studies were conducted in Fischer 344 and Sprague-Dawley (CD) rats at doses up to 1000 mg/kg/day. Doses 250 mg/kg/day caused decreased body weight gain and feed consumption, alterations in red

Hayes’ Handbook of Pesticide Toxicology

blood cells and clinical chemistry parameters (males), perineal soiling, and increased liver weights. Slight centrilobular hepatocellular hypertrophy was observed in high-dose males, while very slight to slight mineralization of the renal pelvic epithelium and very slight to slight hyperplasia of the pelvic epithelium was seen in high-dose females. This observation was consistent with an irritant effect of the mineralized material. Similarly, hyperplasia of the transitional epithelium of the renal pelvis was observed in male and female CD rats at 250 mg/kg/day. The NOAEL was established at 50 mg/kg/day in F344 and 100 mg/kg/day in CD rats (Crissman and Dryzga, 2000; Johnson and Baker, 2000; Stebbins et al., 1998). Similar findings were observed in 4-week and 90-day studies in CD-1 mice. Treatment-related increases in absolute and relative liver weights as well as hepatocellular hypertrophy were observed at doses 100 mg/kg. Based on these findings, the NOAEL was determined to be 10 mg/ kg/day (Crissman and Zablotny, 1998; Yano et al., 2000). In a 4-week study, Beagle dogs received up to 0.90% (highest palatable concentration) in the diet. Body weights and feed consumption were reduced in high-dose males and females. Variable but treatment-related increases in serum enzyme activities (ALT, AST, and AP) were observed at doses  0.45%. Treatment-related increases in liver and thymic weights were also observed and histopathological liver and kidney effects (all dose levels in females) were also noted, and alterations in hematological parameters were observed. A NOAEL was not established in female dogs; it was 29 mg/kg/day in males. Similar effects on the liver and kidneys were observed in a subchronic study in which the NOEL was 0.045% or 18 and 20 mg/kg/day in males and females, respectively (Stebbins and Baker, 1998, 2000). In a 28-day percutaneous toxicity study, there were no treatment-related effects at a limit dose of 1000 mg/kg/day (Stebbins et al., 2000).

87.7.2.3  Chronic Toxicity and Carcinogenicity A combined chronic toxicity and carcinogenicity study was conducted in the Fischer 344 (F344) rat and a second species carcinogenicity study was conducted in the CD-1 mouse. Consistent with the subchronic studies, treatmentrelated histopathologic changes were observed in the kidneys and urinary bladder in chronic toxicity studies. In Fischer 344 rats, chronic exposure to penoxsulam at dosages of 50 and 250 mg/kg/day resulted in an increase in severity of chronic progressive glomerulonephropathy, a common lesion in aging rats. Effects observed at 250 mg/ kg/day included increased hyperplasia of the pelvic epithelium, crystals in the urinary bladder lumen, and hyperplasia of the urinary bladder mucosa. These and similar findings in

Chapter | 87  Toxicology of Triazolopyrimidine Herbicides

other repeat-dose studies are likely attributable to an irritant effect of precipitated penoxsulam, which was identified as the main component of the urinary bladder crystals (Dryzga and Markham, 2006). Body weights were negatively affected at the top dose in both males and females, while liver and kidney weights were increased. The only neoplasm with a statistically identified increase was large granular lymphocytic (LGL) leukemia in male rats. This was considered unrelated to treatment since a dose response was not present, the incidence was within the historical range reported by the National Toxicology Program (NTP), and it is a common tumor of male F344 rats. This opinion was supported by an independent Pathology Working Group review and various agency evaluations of the data, including the U.S. EPA, which accordingly has not performed a quantitative cancer risk assessment for penoxsulam. Penoxsulam was noncarcinogenic in an 18-month dietary study in CD1 mice; there were no adverse effects at the highest dose tested (Hardisty, 2002; Johnson et al., 2002; U.S. EPA, 2009; Yano and Day, 2002).

87.7.2.4  Mutagenicity In a battery of tests, penoxsulam showed no evidence of genotoxic potential. These tests included an in vitro bacterial reverse mutation assay (Ames test), an in vitro forward mutation test in Chinese hamster ovary cells (CHO/HGPRT assay), a mammalian cytogenetics test in rat lymphocytes, and an in vivo cytogenetic assay in mouse bone marrow cells (micronucleus test) (Day and Shabrang, 1999; Lawlor, 1999; Linscombe et al., 1999a,b).

87.7.2.5  Neurotoxicity Acute and chronic (1-year) neurotoxicity studies were conducted in Fischer 344 rats and did not reveal any specific neurotoxic effect. The NOEL in the acute study was 2000 mg/kg/day, while it was 250 mg/kg/day in the chronic study (Marable et al., 2002; Spencer and Johnson, 2000).

87.7.2.6  Reproductive Toxicity In oral gavage developmental toxicity studies, there were no effects on intrauterine development or prenatal survival at doses as high as 1000 mg/kg/day in CD rats and 75 mg/ kg/day in New Zealand White rabbits. Some maternal body weight effects and clinical signs were seen at the high-dose levels. In rats, the embryo-fetal NOEL was 1000 mg/kg (highest dose tested), while the maternal NOEL was 500 mg/kg/day. In rabbits, the NOAEL for embryo-fetal effects was 75 mg/kg/day (highest dose tested), while the maternal NOAEL was 25 mg/kg/day. In a multigeneration reproduction study in CD rats, dosages ranged from 30 to 300 mg/kg/day. There were no effects on reproduction at the highest dose tested (reproductive NOEL  300 mg/kg/day).

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The parental NOEL was 30 mg/kg/day based on hyperplasia of the renal pelvic epithelium at 100 mg/kg/day. The NOEL for offspring was also 30 mg/kg/day based on slight delays in preputial separation likely secondary to decreased body weight (Carney et al., 2000, 2002; Marty et al., 2001).

87.7.2.7  Absorption, Distribution, Metabolism, and Excretion In metabolism studies in F344 rats, a single oral gavage dose of 5 mg/kg of 14C-penoxsulam was readily and extensively absorbed and rapidly eliminated. Saturation of absorption was evident at a higher dose of 250 mg/kg/day. At the low dose, excretion of radiolabeled penoxsulam was primarily through the fecal route in males, while females eliminated radioactivity primarily in the urine and to a lesser extent in the feces. At the higher dose, elimination was primarily through the feces for both genders, likely representing saturation of absorption. The majority of the radioactivity (86% in urine) was excreted as unchanged penoxsulam; however, penoxsulam is biotransformed to 36 different metabolites. Metabolites identified in excreta were generally detected at low levels, although some fecal metabolites were shown to account for 5% of the administered dose. Penoxsulam was shown to be metabolized primarily through demethylation and ring hydroxylation with subsequent conjugation to glucuronic acid or glutathione. The rapid absorption and excretion of penoxsulam indicate no potential for bioaccumulation. Absorption at 24 h following dermal exposure of F344 rats was approximately 2% for undiluted and 0.4% for a spray dilution (0.03 g/l) of penoxsulam (Mandrala et al., 2002).

87.7.3  Toxicity to Humans No studies are available on intentional human exposure. Risk assessment calculations for the general population and for pesticide handlers indicate a low-risk estimate. Maximum residue levels (MRLs) of 0.01–0.02 ppm for crops have been established based on the analytical limit of detection and a reference dose of 0.147 mg/kg/day resulting from the chronic dog study (U.S. EPA, 2007c).

87.8  Pyroxsulam 87.8.1  Identity, Properties, and Uses 87.8.1.1  Chemical Name The IUPAC name for pyroxsulam is N-(5,7-dimethoxy[1,2, 4]triazolo[1,5-a]pyrimidin-2-yl)-2-methoxy-4-(trifluorom ethyl)pyridine-3-sulfonamide; the CAS name is N-(5,7-di methoxy[1,2,4]triazolo[1,5-a]pyrimidin-2-yl)-2-methoxy4-(trifluoromethyl)-3-pyridinesulfonamide.

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87.8.1.2  Structure See Figure 87.1 and Table 87.1.

87.8.1.3  Synonyms Pyroxsulam is also known as BAS 770H, X666742, XR742, XDE-742, DE-742; its CAS number is 422556-08-9. It is sold in plant protection products, either alone or in combination with other herbicide active ingredients, under a variety of registered trade names including: Crusader, Pallas, Perun, and Simplicity herbicides.

87.8.1.4  Physical and Chemical Properties Pyroxsulam has an off-white color. It has a vapor pressure 7.5  1010 Pa at 20°C. The empirical formula is C14H13F3N6O5S and the molecular weight is 434.4. The solubility of pyroxsulam in water at 20°C and pH 7 is 3.20 g/l.

87.8.1.5  Uses Pyroxsulam is a broad-spectrum systemic grass and broadleaf postemergence herbicide.

87.8.2  Toxicity to Laboratory Animals 87.8.2.1  Acute Exposure Pyroxsulam has very low acute toxicity by the oral, dermal, and inhalation routes. It caused slight, transient skin and eye irritation and was identified as a skin sensitizer in guinea pigs by the Magnusson and Kligman maximization method. The oral and dermal LD50 values were each greater than 2000 mg/kg in rats while the 4-h inhalation LC50 was 5.0 mg/L (Gamer and Leibold, 2003a,b, 2004; Kaufmann and Leibold, 2003a,b; Lowe, 2007).

87.8.2.2  Repeated Exposure Short-term dietary studies in rats, mice, and dogs and a short-term dermal study in rats were conducted. The liver was identified as the primary target organ in all species; serum total cholesterol was also universally increased. In a 4-week study in Fischer 344 (F344) rats, there were no adverse effects at doses up to 1150 mg/kg bw/ day (Stebbins and Day, 2001). In a 13-week study, body weight gain was reduced in females at 1000 mg/kg bw/day. There was evidence for a slight treatment-related effect on the liver of males at 1000 mg/kg bw/day: increased relative liver weight (and serum cholesterol) but without associated histopathological change it was not considered an adverse finding. By 28 days post-treatment, there was complete recovery of liver weight and a marked recovery of the cholesterol level (Stebbins et al., 2003). Similarly,

Hayes’ Handbook of Pesticide Toxicology

in a 13-week study in CD-1 mice the only effects were increased liver weight and serum cholesterol at a limit dose of 1000 mg/kg bw/day (Johnson et al., 2003). In a 13-week study in Beagle dogs, apparent effects comprised reduced body weight gain, increased liver weight, hepatocellular hypertrophy, and increased serum cholesterol at a high dose of 884 and 1142 mg/kg bw/day in males and females, respectively (Stebbins and Baker, 2003). In a 1-year study, increased liver weight without hepatocyte hypertrophy was seen in both sexes at the top dose of approximately 600 mg/kg bw/day. There was also evidence of increased serum cholesterol and alkaline phosphatase activity, which might reflect minimal cholestasis (Stebbins and Dryzga, 2004). Based on these studies, the overall short-term NOAEL was 90 mg/kg bw/day based on the minor effects seen at high-dose levels in the range of approximately 600–1000 mg/kg bw/day. In a 14-day dermal toxicity study in F344 rats, there were no treatment-related effects at the limit dose of 1000 mg/kg bw/day (Kaspers, 2004).

87.8.2.3  Chronic Toxicity and Carcinogenicity A combined chronic toxicity and carcinogenicity study was conducted in the Fischer 344 (F344) rat and a second species carcinogenicity study was conducted in the CD-1 mouse. In the F344 rat study, dose levels of 0, 10, 100, and 1000 mg/kg bw/day were administered via the diet for 2 years (Stebbins and Brooks, 2005). Body weight gain and feed consumption were slightly reduced in high-doselevel females. Slight increases in serum total cholesterol and liver weight, in the absence of any other associated changes including histopathologically, were considered nonadverse, adaptive changes. There were no treatmentrelated increases in neoplasms in male or female rats at any dose level, indicating that pyroxsulam did not have an carcinogenic potential under the conditions of this study up to and including a limit dose. The CD-1 mouse study dose levels of 0, 10, 100 and 1000 mg/kg bw/day were administered via the diet for 18 months (Johnson et al., 2005). The only treatmentrelated effects were in the liver of high-dose males and comprised increased liver weight, increased incidence of foci of altered hepatocytes, and a possible marginal increased incidence and number of hepatocellular adenomas and carcinomas. Doubt exists over an association with treatment as there was no dose–response relationship and differences from concurrent controls were not statistically identified but values for adenomas were just outside historical control ranges. Toxicokinetic data demonstrated systemic AUC values were less than dose proportional at the high-dose level but increased 30-fold from 10 to 1000 mg/kg bw. Pyroxsulam did not to cause any ­genotoxicity effects in a

Chapter | 87  Toxicology of Triazolopyrimidine Herbicides

complete battery of tests, including an in vivo mouse liver UDS assay conducted with a limit dose of 2000 mg/kg bw. Under the conditions of this study, the NOAEL was 100 and 1000 mg//kg bw/day for male and female mice, respectively.

87.8.2.4  Mutagenicity In a battery of tests, pyroxsulam showed no evidence of genotoxic potential. These tests comprised an in vitro bacterial reverse mutation assay (Ames test; Engelhardt and Leibold, 2003), an in vitro mammalian cytogenetics test in CD rat lymphocytes (Schistler, 2006), an in vitro mammalian forward mutation test in Chinese hamster ovary cells (CHO/HGPRT assay; Schisler and Grundy, 2006), and two in vivo assays in the CD-1 mouse: a cytogenetic assay in bone marrow cells (micronucleus test; Spencer and Grundy, 2004), and an unscheduled DNA synthesis (UDS) assay (Beevers, 2006), both of which involved oral gavage dosing at the limit dose of 2000 mg/kg bw.

87.8.2.5  Reproductive Toxicity Developmental studies in CD rats and NZW rabbits and a two-generation reproduction study have been completed for pyroxsulam. In the reproduction study in the CD rat, dose levels of 0, 10, 100, and 1000 mg/kg bw/day were administered via the diet (Carney et al., 2005). There was no evidence of systemic toxicity or adverse effects on any parameter of reproductive function of the parental animals, nor on survival, growth, or development of F1 or F2 offspring. The rat developmental toxicity study assessing dose levels of 0, 100, 300, and 1000 mg/kg bw/day via gavage produced no treatment-related maternal toxicity and no indications of embryo-fetal toxicity or teratogenicity (Carney and Tornesi, 2005). In the rabbit, doses of 0, 30, 100, and 300 mg/kg bw/day were also administered by gavage (Sloter, 2005). The high-dose level was based on body weight effects at 300, 600, and 1000 mg/kg bw/day in a probe dose-range finding study. However, no maternal toxicity, or effects on embryo-fetal survival and development, occurred in the main study.

87.8.2.6  Absorption, Distribution, Metabolism, and Excretion Studies of absorption, distribution, metabolism, and excretion (ADME) have been conducted in rats and mice. The study in male Fischer 344 (F344) rats was of typical regulatory test guideline design and investigated two dose levels, a low and a high dose of 10 and 1000 mg/kg body weight, respectively (Hansen et al., 2005). The mouse study was conducted to provide toxicokinetic information to aid investigation of the possible mechanism and relevance

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for human risk assessment of apparent treatment-related liver tumors in high-dose-level male mice following lifetime exposure to pyroxsulam (Hansen et al., 2006). The study investigated the three dose levels used in the mouse carcinogenicity study (10, 100, and 1000 mg/kg bw) in males; limited data were also generated for females. In rats, a single oral gavage dose of 10 mg/kg of 14Cpyroxsulam was readily and extensively absorbed (ca. 75%) and rapidly eliminated. Absorption was slightly lower at the limit dose of 1000 mg/kg. Excretion of radiolabeled pyroxsulam was rapid and virtually complete within 48 hours, primarily through the urine but with a small biliary component (ca. 17% of the total) identified from animals dosed by the intravenous route. The majority of the radioactivity (85%) was excreted as unchanged pyroxsulam and only one substance, the 2-desmethyl metabolite, representing at least 5% of the administered dose. There were no differences in the findings between single or repeat, 15-day, dosing and no evidence for metabolic induction (no alteration in metabolism of pyroxsulam). The rapid and extensive excretion with very low levels in carcass at 48 h after dosing (1% of administered dose) indicate no potential for accumulation. The investigatory study in the mouse revealed results similar to those outlined for the rat. Absorption was rapid and extensive and plasma elimination was rapid (t½ of 2–3 h) and excretion was primarily in urine.

87.8.3  Toxicity to Humans No studies are available on intentional human exposure. Risk assessment calculations for the general population and for pesticide handlers indicate a low-risk estimate. A reference dose of 1 mg/kg/day was established on the basis of a NOAEL of 100 mg/kg/day from the carcinogenicity study in mice. Dietary risk evaluation assuming 100% of crop are treated and residues are at the established tolerance level (0.01 ppm) indicates that the chronic dietary risk estimates for the U.S. population and all population subgroups utilize 0.1% of the RfD (U.S. EPA, 2007b). Risk assessment calculations for the general population and for pesticide operators, field workers, and bystanders indicate exposures significantly below proposed reference doses for pyroxsulam.

Conclusion In conclusion, the triazolopyrimides are a structurally-related class of herbicides which act to inhibit the enzyme acetolactate synthase (ALS) in plants. This target does not exist in animals and accordingly this class of chemicals exhibits low toxicity in mammals. The triazolopyrimidine herbicides have been comprehensively evaluated in guideline and GLP compliant toxicity studies required for the registration and

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authorization of pesticides in various geographies throughout the world. In general, they exhibit very low mammalian toxicity as assessed through acute, short-term, long-term (chronic), genotoxicity, reproduction, developmental, and neurotoxicity studies. In repeat-dose toxicity studies, the liver and kidneys have been identified as target organs with effects that were often adaptive in nature generally observed only at excessively high-dose levels. In addition, the triazolopyrimidines were shown to be rapidly absorbed and excreted, have a low potential for bioaccumulation, and in general are not extensively metabolized.

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Cosse, P. F., and Berdasco, N. M. (1992d). Dow AgroSciences LLC, unpublished data. Cosse, P. F., Yano, B. L., and Stott, W. T. (1989). Dow AgroSciences LLC, unpublished data. Crissman, J. W., and Dryzga, M. D. (2000). Dow AgroSciences LLC, unpublished data. Crissman, J. W., and Zablotny, C. L. (1998). Dow AgroSciences LLC, unpublished data. Day, S. J., and Shabrang, S. N. (1999). Dow AgroSciences LLC, unpublished data. DeFronzo, R. A. (1980). Hyperkalemia and hyporeninemic hypoaldosteronism. Kidney Int. 17, 118–134. Dietrich, D. R., and Swenberg, J. A. (1991). Preneoplastic lesions in rodent kidney induced spontaneously or by non-genotoxic agents: Predictive nature and comparison to lesions induced by genotoxic carcinogens. Mutat. Res. 248, 239–269. Domoradzki, J. Y., Stewart, H. S., Hansen, S. C., Brzak, K. A., and Dryzga, M. D. (1995). Dow AgroSciences LLC, unpublished data. Dryzga, M. D., and Markham, D. A. (2006). Dow AgroSciences LLC, unpublished data. Dryzga, M. D., Johnson, K. A., and Cieslak, F. S. (1996). Dow AgroSciences LLC, unpublished data. Eiam-ong, S., Laski, M. E., Kurtzman, N. A., and Sabatini, S. (1994). Effect of respiratory acidosis and respiratory alkalosis on renal transport enzymes. Am. J. Physiol. 267, F390–F399. Engelhardt, G., and Leibold, E. (2003). Dow AgroSciences LLC, unpublished data. Environmental Protection Agency (EPA) (1993). Pesticide tolerance for flumetsulam. 40 CFR 180. Federal Register 58(207), 57966 (Thursday, October 28, 1993). Environmental Protection Agency (EPA). (1997a). Cloransulam-methyl: Pesticide fact sheet. OPPTS 7501C (available at http://www.epa.gov/ opprd001/factsheets/cloransu.html). Environmental Protection Agency (EPA). (1997b). Cloransulam-methyl: Pesticide tolerances. 40 CR 180, Federal Register 62(182), 49158 (Friday, September 19, 1997) (available at http://www.epa.gov/fedrgstr/EPA-PEST/1997/September/Day-19/p24939.htm). Environmental Protection Agency (EPA). (1998). Diclosulam: Notice of filing of pesticide petitions. 40 CR 180, Federal Register 63(224), 64484 (Friday, November 20, 1998) (available at http://www.epa.gov/ fedrgstr/EPA-PEST/1998/November/Day-20/p31066.htm). Environmental Protection Agency (EPA). (2007a). Florasulam: Human Health Risk Assessment for Proposed Use on Cereal Grains (Wheat, Oats, Barley, Rye, and Triticale). May 31, 2007. Environmental Protection Agency (EPA). (2007b). Pyroxsulam Human health Risk Assessment for Proposed Uses on Wheat. (December 19, 2007). Environmental Protection Agency (EPA). (2007c). Penoxsulam. Human Health Risk Assessment for Proposed Uses on Fish and Shellfish. (June 18, 2007). Environmental Protection Agency (EPA). (2008). Flumetsulam. Amended Human Health Assessment Scoping Document in Support of Registration Review (August 12, 2008). Environmental Protection Agency (EPA). (2009) Penoxsulam. Pesticide Tolerances: Final Rule (April 14, 2009). Gamer, A. O., and Leibold, E. (2003a). Dow AgroSciences LLC, unpublished data. Gamer, A. O., and Leibold, E. (2003b). Dow AgroSciences LLC, unpublished data.

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Quast, J. F., Haut, K., and Kociba, R. J. (1997). Dow AgroSciences LLC, unpublished data. Redmond, J. M., and Johnson, K. A. (1996a). Dow AgroSciences LLC, unpublished data. Redmond, J. M., and Johnson, K. A. (1996b). Dow AgroSciences LLC, unpublished data. Redmond, J. M., and Kociba, R. J. (1996). Dow AgroSciences LLC, unpublished data. Schisler, M. R. (2006). Dow AgroSciences LLC, unpublished data. Schisler, M. R., and Grundy, J. (2006). Dow AgroSciences LLC, unpublished data. Scortichini, B. H., and Kociba, R. J. (1997). Dow AgroSciences LLC, unpublished data. Shankar, M. R., and Johnson, K. A. (1996). Dow AgroSciences LLC, unpublished data. Shankar, M. R., Stebbins, K. E., and Haut, K. T. (1993). Dow AgroSciences LLC, unpublished data. Spencer, P. J., and Johnson, K. A. (2000). Dow AgroSciences LLC, unpublished data. Spencer, P. J., Stebbins, K. E., and Albee, R. R. (1995). Dow AgroSciences LLC, unpublished data. Spencer, P. J., and Grundy, J. (2004). Dow AgroSciences LLC, unpublished data. Stebbins, K. E., and Baker, P. C. (1998). Dow AgroSciences LLC, unpublished data. Stebbins, K. E., and Baker, P. C. (2000). Dow AgroSciences LLC, unpublished data. Stebbins, K. E., and Brooks, K. J. (2005). Dow AgroSciences LLC, unpublished data. Stebbins, K. E., and Haut, K. T. (1993). Dow AgroSciences LLC, unpublished data. Stebbins, K. E., and Haut, K. T. (1994). Dow AgroSciences LLC, unpublished data. Stebbins, K. E., and Haut, K. T. (1997). Dow AgroSciences LLC, unpublished data. Stebbins, K. E., Zielke, G. J., and Hanley, T. R. (1990). Dow AgroSciences LLC, unpublished data. Stebbins, K. E., Haut, K. T., and Stott, W. T. (1996). Dow AgroSciences LLC, unpublished data. Stebbins, K. E., Day, S. J., and Cieszlak, F. S. (1998). Dow AgroSciences LLC, unpublished data. Stebbins, K. E., Yano, B. L., and Baker, P. C. (2000). Dow AgroSciences LLC, unpublished data. Stewart, H. S., Yano, B. L., Vedula, U., and Stott, W. T. (1992). Dow AgroSciences LLC, unpublished data. Stewart, H. S., Yano, B. L., and Stott, W. T. (1993). Dow AgroSciences LLC, unpublished data. Stewart, H. S., McNett, D. A., Timchalk, C., and Brzak, K. A. (1996). Dow AgroSciences LLC, unpublished data. Stokes, J. B. (1993). Ion transport by the collecting duct. Semin. Nephrol. 13, 202–212. Stott, W. T., Quast, J. F., Cieszlak, F. S., Schuetz, D. J., and Phillips, J. E. (1986). Dow AgroSciences LLC, unpubliahed data. Stott, W. T., Yano, B. L., Beyer, J. E., Kropscott, B. E., and Eddy, S. L. (1991). Dow AgroSciences LLC, unpubliahed data. Sullivan, J. M., and Cronin-Singleton, N. (1995). Dow AgroSciences LLC, unpubliahed data. Sullivan, J. M., and Singleton, N. (1995). Dow AgroSciences LLC, unpublished data.

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Swaim, L. D., and Szabo, J. R. (1992). Dow AgroSciences LLC, unpublished data. Szabo, J. R., and Davis, N. L. (1992). Dow AgroSciences LLC, unpublished data. Szabo, J. R., and Davis, N. L. (1993a). Dow AgroSciences LLC, unpublished data. Szabo, J. R., and Davis, N. L. (1993b). Dow AgroSciences LLC, unpublished data. Szabo, J. R., and Davis, N. L. (1994). Dow AgroSciences LLC, unpublished data. Szabo, J. R., and Rachunek, B. L. (1992). Dow AgroSciences LLC, unpublished data. Szabo, J. R., Davis, N. L., and Campbell, R. A. (1992). Dow AgroSciences LLC, unpublished data. Timchalk, C., Dryzga, M. D., and Kropscott, B. (1988). Dow AgroSciences LLC, unpublished data. Timchalk, C., Dryzga, M. D., Johnson, K. A., Eddy, S. L., Freshour, N. L., Kropscott, B. E., and Nolan, R. J. (1996). Comparative pharmacokinetics of [14C]metosulam (N[2,6-dichloro-3-methylphenyl]-5, 7-dimethoxy-1,2,4-triazolo-[1,5a]-pyrimidine-2-sulfonamide) in rats, mice and dogs. J. Appl. Toxicol. 17, 9–21. Tsuruoka, S., and Schwartz, G. J. (1996a). Adaptation of rabbit cortical collecting duct HCO–3 transport to metabolic acidosis in vitro. J. Clin. Invest. 97, 1076–1084. Tsuruoka, S., and Schwartz, G. J. (1996b). Metabolic acidosis stimulates H secretion in the perfused rabbit outer collecting duct of the inner stripe. J. Am. Soc. Nephrol. 7, 1262. UK Ministry of Agriculture, Fisheries, and Food. (1996). “Evaluation of Fully Approved or Provisionally Approved Products. Evaluation on: Metosulam.” Pesticide Safety Directorate, Issue 13, No. 148, Ministry of Agriculture, Fisheries, and Food. Vedula, U., Stebbins, K. E., and Breslin, W. J. (1992). Dow AgroSciences LLC, unpublished data. Verlander, J. W., Madsen, K. M., and Tisher, C. C. (1991). Structural and functional features of proton and bicarbonate transport in the rat collecting duct. Semin. Nephrol. 11, 465–477. Verlander, J. W., Madsen, K. M., Cannon, J. K., and Tisher, C. C. (1994). Activation of acid-secreting intercalated cells in rabbit collecting duct with ammonium chloride loading. Am. J. Physiol. 266, F633–F645. Walker, M. D. (1996). Dow AgroSciences LLC, unpublished data. Weiner, I. D. (1997). Dow AgroSciences LLC, unpublished data. Weiner, I. D., and Wingo, C. S. (1997). Hypokalemia—consequences, causes and correction. J. Am. Soc. Nephrol. 8, 1179–1188. Wingo, C. S., and Cain, B. D. (1993). The renal H-K-ATPase: Physiological significance and role in potassium homeostasis. Ann. Rev. Physiol. 55, 323–347. Yano, B. L., and Day, S. J. (2002). Dow AgroSciences LLC, unpublished data. Yano, B. L., Firchau, H. M., and Stott, W. T. (1987). Dow AgroSciences LLC, unpublished data. Yano, B. L., Firchau, H. M., and Quast, J. F. (1988). Dow AgroSciences LLC, unpublished data. Yano, B. L., Cosse, P. F., and Corley, R. A. (1991). Dow AgroSciences LLC, unpublished data. Yano, B. L., Cieszlak, F. S., and Day, S. J. (2000). Dow AgroSciences LLC, unpublished data. Zablotny, C. L. (1996). Dow AgroSciences LLC, unpublished data. Zablotny, C. L., and Carney, E. W. (1997). Dow AgroSciences LLC, unpublished data.

Chapter | 87  Toxicology of Triazolopyrimidine Herbicides

Zablotny, C. L., Stebbins, K. E., and Breslin, W. J. (1993). Dow AgroSciences LLC, unpublished data. Zablotny, C. L., Breslin, W. J., Stebbins, K. E., and Engle, K. E. (1994). Dow AgroSciences LLC, unpublished data. Zablotny, C. L., Breslin, W. J., and Quast, J. F. (1996). Dow AgroSciences LLC, unpublished data.

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Zempel, J. A., Grandjean, M., and Szabo, J. R. (1988). Dow AgroSciences LLC, unpublished data. Zempel, J. A., Mensik, D. C., and Szabo, J. R. (1990). Dow AgroSciences LLC, unpublished data. Zielke, G. J., Hanley, T. R., and Yano, B. L. (1988). Dow AgroSciences LLC, unpublished data.

Section XIII

Fungicides

(c) 2011 Elsevier Inc. All Rights Reserved.

Chapter 88

A Toxicological Assessment of Sulfur as a Pesticide Derek W. Gammon, Thomas B. Moore and Michael A. O’Malley Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, CA, USA

88.1  Introduction 88.1.1  Usage Elemental sulfur is the most heavily used crop protection chemical in California (Table 88.1) as well as in the United States. In 1993–1995, for example, annual usage was about 70 million pounds active ingredient (a.i.) in California, which is about one-third of the total weight of pesticides used in agriculture. It is generally applied to crops as a dust to combat fungal disease, at rates of approximately 10– 30 lbs per application per acre, as well as being used for postharvest disease control. The range of fungal diseases controlled by sulfur includes brown rot, scab, mildew, powdery mildew, leafspot, and rusts (Farm Chemicals Handbook, 1998). It is also used, to a lesser extent, for the control of mites and insects (fleahoppers), which may be secondary to fungal damage of the plant. Multiple applications are often needed for crops which are particularly susceptible to fungal attack. The main crops on which sulfur is used in California (1995) are grapes (71%), tomatoes (12%), and sugar beet (8%). It has become an important component of integrated pest management (IPM) systems since it can be used in “organic” farming.

88.1.2  Environmental Fate As a natural substance, environmental fate requirements for sulfur in the United States have been waived by U.S. EPA. In a variety of literature reports, sulfur has been shown to be oxidized, in the presence of water and soil, to the sulfite and then the sulfate, i.e., sulfuric acid. The supplementation Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

of fertilizers with sulfur has been used intentionally to acidify soils which are too alkaline for a particular crop. Conversely, the use of sulfur as a fungicide can make the soil too acidic for the continued optimal growth of a particular crop. For example, in Southern Tanzania, sulfur dust was used to control powdery mildew on cashew nut trees. After 4 years, the topsoil pH was reduced by 0.7 units, to below pH 5.5, the ideal pH for cashew nut tree yield (Majule et al., 1997). However, in the case of a 3-year field study on highbush blueberry bushes in mineral soil, sulfur amendment increased both early growth and blueberry yield. It was concluded that the effects of sulfur were probably mediated by a decrease in soil pH with corresponding increases in Mn and Fe levels (Haynes and Swift, 1986). It was also found, in a lysimeter study, that increasing the sulfur content of soil led to a rise in sulfur content of plants, such as corn, wheat, barley, sunflower, and mustard (Gador and MotowickaTerelak, 1986). Sulfur is, of course, a natural constituent of plants, as it is of all organisms. In addition to elemental sulfur, which may be present from pesticidal exposure, sulfur is commonly found as sulfates and in the amino acids cysteine, methionine, and glutathione. Agricultural practice may introduce elevated sulfur levels into arable land inadvertently by, for example, the use of (animal-derived) manure or other fertilizers as well as from the use of pesticides. By far the greatest contribution in the latter case is likely to be from elemental sulfur used as a fungicide. As described below, there are an increasing number of instances of dermatitis in farm workers and, in many ways more serious, an elevated number of cases of disease in ruminants caused by exposure to high levels of sulfur, and several case reports are given.

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Table 88.1  Usage of Elemental Sulfur (a.i.) in California, 1993–1995a,b Crop

1994c in thousands

1995 in thousands

1993 in thousands

lbs.

Acresd

lbs.

Acresd

lbs.

Acresd

Alfalfa

294

13.2

227

10.7

139

6.4

Almonds

184

42.4

126

26

48.4

11.0

Bermuda grass

162

5.9

31.2

1.26

62.6

2.2

Cantaloupe

634

31.3

357

17.9

363

20.9

Carrots

177

10.8

347

15.2

454

17.9

Cotton

460

17.3

240

10.9

319

13.3

Date

741

11.3

655

10.9

838

14.1

Grapes, table

26,200

2970

24,900

2660

25,600

2630

Grapes, wine

23,200

2460

23,500

2340

25,700

2450

Lemons

193

4.8

190

5.8

168

5.3

Melons

133

5.8

282

11.8

349

17.0

Nectarines

277

35.3

240

31.0

198

25.5

Peaches

873

83.3

644

69.8

588

66.1

Pears

170

13.6

169

12.9

197

16.2

Peas

204

10.1

305

14.5

281

14.6

Pistachio

500

39.4

592

30.7

846

33.9

Plums

259

25.0

262

23.7

267

26.9

Prunes

224

23.0

135

13.8

157

17.4

Strawberry

238

64.7

262

61.9

256

63.9

Sugar beet

5300

171

6760

210

6870

218

Tomatoes, fresh

1270

62.8

1620

76.9

1380

70.7

Tomatoes, processing

7250

292

7810

290

7700

219

TotaL (lbs.)

69.8 million

70.5 million

73.5 million

a

DPR, 1995, 1996a, 1996b. All crops which received more than 100,000 lbs. a.i. in 1995 are included. c 132,000 lbs. of sulfur were applied to 8600 acres of oranges in 1994. d Acres treated include multiple applications to the same land. b

Another source of soil acidity caused by sulfur is from industrial pollution. Typically from the burning of coal or oil, sulfur dioxide or hydrogen sulfide can be liberated into the atmosphere. This gaseous sulfur can return to earth following rainfall, producing the so-called acid rain phenomenon since sulfuric acid can readily be produced from soil oxidation of sulfur.

88.2  Toxicology profile of elemental sulfur Acute toxicity categories for products containing elemental sulfur were assigned according to FIFRA Pesticide Assessment Guidelines, Subdivision F, Hazard

Evaluation: Human and Domestic Animals, Revised Ed., Nov., 1984. The results are summarized in Table 88.2.

88.2.1  Acute Exposure Oral Toxicity: 81-1 A single-dose limit test was conducted using the rat (five/ sex), dosed by gavage at 5000 mg/kg. Lower doses are only required if there is 20% mortality per sex. Seventeen of 20 formulations had a LD50   5000 mg/kg (Category IV); three were between 500 and 5000 mg/kg (Category III).

88.2.2  Acute Exposure Dermal Toxicity: 81-2 A single-dose limit test was conducted using the rabbit (five/sex), dosed on the skin at 2000 or 5000 mg/kg. Lower

Chapter | 88  A Toxicological Assessment of Sulfur as a Pesticide

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Table 88.2  Acute Toxicity of Sulfur Formulations Used in California Sulfur product

Orala

Dermalb

Inhal.c

Eye irrit.d

Dermal irrit.e

Dermal sensit.f

Special electric refined super adhesive dusting (98% a.i.)

IV

III

—

III

IV

—

Manufacturing use (98%)

IV

III

IV

III

IV

—

Spray sulfur (98%)

IV

III

IV

III

IV

—

Valor Brands Products Dusting Sulfur (98%)

IV

III

IV

III

IV

0

90% Sulfur WP (90%)

IV

III

IV

III

IV

0

Bensul 85 (85%)

IV

III

IV

III

IV

0

Clean Crop Apple & peach Koloform Fungicide (84%)

IV

III

IV

III

IV

0

LX 112–2 (53%)

IV

III

IV

III

IV

—

Sulfur 6L (51%)

IV

III

IV

III

IV

0

Happy Jack Sarcoptic Mange Medicine (28%)

IV

III

—

III

III

—

Safer Garden Fungicide Concentrate (12%)

IV

III

IV

III

IV

0

Formula 242 (0.4%)

IV

—

—

IV

IV

—

XF-97097 (96.75%), with myclobutanil, 0.5%

IV

IV

IV

III

IV

—

BT 320 Sulfur 50 (50%), with BT, 0.064%

IV

III

IV

III

IV

0

Britz BT50 & Sulfur Dust (50%) BT, 0.064%

IV

III

IV

III

IV

—

Cook/Sevin Plus Multi-purpose Garden Dust (30%) with carbaryl 5%, PBO 0.45%, permethrin 0.03%

III

III

III

IV

IV

—

Britz Botran 6–25 Dust (25%) with dichloran, 6%

IV

III

IV

III

IV

—

Britz Copper Sulfur Dust (25%) 15–25 with Cu, 15%

IV

III

IV

II

IV

—

Copper/Sulfur Flowable (15.5%) with Cu sulfate, 27.5%

III

III

III

II

IV

—

Kocide 404S (15%) with Cu hydroxide, 26%

III

III

III

I

IV

—

a

Category IV: LD50   5000 mg/kg; Category III: LD50  500–5000 mg/kg. Category IV: LD50   5000 mg/kg; Category III: LD50  2000–5000 mg/kg. c Category IV: LC50   2 mg/l; Category III: LC50   0.5–2 mg/l. d Category IV: minimal effects, clearing in 24 h; Category III: corneal involvement or irritation, clearing in 7 days; Category II: corneal involvement or irritation, clearing in 8–21 days; Category I: corrosive (irreversible ocular damage) or corneal involvement or irritation, clearing in 21 days. e Category IV: mild or slight irritation (no irritation or slight erythema); Category III: moderate irritation at 72 hours (moderate erythema). f Buehler test: score of 0 (no erythema) to 3 (severe erythema, with or without edema). b

doses are only required if there is 20% mortality per sex. There were no compound-related, acute mortalities for any of these formulations at the tested doses of 2000 or 5000 mg/kg. This suggests that it may be more appropriate to consider them all in Category IV rather than III.

88.2.3  Acute Exposure Inhalation   Toxicity: 81-3 A single-dose limit test was conducted using the rat (five/ sex), dosed by inhalation at 2 mg/l for 4 h. Lower doses are only required if there is 20% mortality per sex. Fourteen of 17 formulations had a LC50  2 mg/l (Category V); three were between 0.5 and 2 mg/l (Category III).

88.2.4  Primary Eye Irritation: 81-4 A single-dose limit test was conducted using the rabbit (six of either sex), dosed in one eye at 0.1 ml/animal (or 100 mg for a solid). The other eye served as the untreated control and responses were graded (Table 88.2). Two of 20 formulations showed minimal effects, clearing within 24 h (Category IV); 15 showed corneal involvement or irritation, clearing in 7 days (Category III); two showed corneal involvement or irritation, clearing in 8–21 days (Category II); one showed corrosive (irreversible ocular damage) or corneal involvement or irritation, clearing in 21 days (Category I), probably as a result of the copper hydroxide in this formulation.

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88.2.5  Primary Dermal Irritation: 81-5 A single-dose limit test was conducted using the rabbit (six of either sex), dosed at 0.5 ml/in.2 (or 0.5 g/in.2) for 4 h. The responses were graded. Nineteen of 20 formulations caused mild or slight irritation (Category IV); one showed moderate irritation at 72 h (Category III).

88.2.6  Primary Dermal Sensitization: 81-6 Using the Buehler test, induction doses were applied to clipped skin at 0.4 ml or 500 mg/guinea pig (ca. 1–2 g/kg) three times, on a weekly basis, followed 2 weeks later by a challenge dose, to a naive site. Dermal sensitization was measured, as erythema with or without edema, in response to the challenge dose, at 24 and 48 h, on a scale of 0 to 3. All of the (seven) formulations tested scored zero, i.e., were negative.

88.3  Toxicology of sulfur dioxide Sulfur dioxide (SO2) is used as a fumigant because of its antimicrobial properties. It is a colorless gas with a high water solubility. In solution, it hydrates to sulfurous acid (H2SO3), which dissociates in turn to form bisulfite ( HSO 3 ) and sulfite (SO2–3) ions. The bisulfite ion is quite reactive by means of ionic and free radical mechanisms (Shapiro, 1977). Sulfur dioxide is used in California for the treatment of grapes held in cold storage to control the fungus Botrytis cinerea. The recommended treatment rate is up to a 1% gas concentration for up to 20 treatments with 7- to 10-day intervals between treatments depending on the variety of grape. The main crop uses of SO2 in California are summarized in Table 88.3. Sulfur dioxide is used in the United States as a food additive under the authority of the Food and Drug Administration in beer, wine, fruits and vegetables, fruit juices, syrups, meats, and fish. It acts as a preservative in these foods by

Table 88.3  Usage of Sulfur Dioxide (a.i.) in California, 1993–1995a Crop

1995 (lbs.) 1994 (lbs.) 1993 (lbs.)

Grapes, table

144,000

267,000

194,000

Grapes, wine

24,000

1100

12,000

Commodity fumigation

14,000

11,700

48,100

Other fumigation

4000

5500

6000

Structural pests

13,400

—

15,100

Total (lbs.)

200,000

285,000

276,000

a

DPR, 1995, 1996a, 1996b.

being both an antimicrobial and an inhibitor of the enzymes that contribute to the discoloration process. Sulfur dioxide has been used in wine making to selectively inhibit the growth of acetic acid and lactic acid producing bacteria. Product registrations for sulfur dioxide in California are for the 100% compressed gas. Precautionary labeling for these products requires the signal word “Danger” with the wording “inhalation may be fatal or cause serious illness. Prolonged or repeated exposure may cause impaired lung function…. Liquid or excessive vapor exposure can cause serious skin and eye injury. Harmful if swallowed.” When sulfur dioxide is used as a fumigant, respiratory protection is required unless the ambient concentration is less than 2 ppm, which is the threshold limit value for occupational exposure. Case studies of cats and dogs fed fresh pet food preserved with sulfur dioxide resulted in examples of animals suffering from thiamine deficiency (Studdert and Labuc, 1991). These animals demonstrated a syndrome of depression, pupillary dilation, and ataxia, which occasionally progressed to seizures and sudden death caused by acute cardiac failure. In the preserved food samples in which the SO2 content was greater than 800 mg/kg, the thiamine levels were decidedly reduced. In the presence of sulfiting agents such as sulfur dioxide, thiamine is cleaved into its constituent pyrimidine and thiazole moieties, rendering it inactive. It should be noted that the principal toxic effect of elemental sulfur on the central nervous system (CNS) of ruminants is a direct effect of sulfur and not a secondary effect arising from thiamine deficiency. Other investigators examined pigs fed barley with high moisture content that had been treated with sulfur dioxide (Gibson et al., 1987). Treatment of the barley (1% sulfur dioxide (wt/wt)) demonstrated an enhanced preservation of the barley with a significant time delay before mold growth became evident. However, the thiamine content in the barley was greatly reduced, resulting in a thiamine content in the meat of the treated pigs that was 7.6% that of the control animals. These animals gave evidence of cardiac hypertrophy along with reduced feed intake and body weight gain. Once again, it is possible that direct effects of SO2 contributed to the toxicity of the barley to the pigs, rather than these being purely secondary consequences of thiamine deficiency. The World Health Organization specifically recommends that foods that are significant sources of thiamine in the human diet should not be treated with sulfur dioxide or other sulfiting agents. Pollution of the environment has been a major health concern as a consequence of excessive exposures to sulfur dioxide and smoke, for example in the Meuse Valley of Belgium in 1936, in Donora, Pennsylvania in 1948, and in London in 1952. In London, where 4000 deaths and numerous incidences of illness were attributed to the exposure, atmospheric sulfur dioxide levels achieved a daily average as high as 1.34 ppm. Pulmonary effects manifested by exposure to sulfur dioxide are attributable to its irritancy.

Chapter | 88  A Toxicological Assessment of Sulfur as a Pesticide

Exposure to sulfur dioxide alone results in direct effects on the nasopharynges and trachea with reduced transport of the mucous layer either due to cessation of ciliary movement in an acute exposure or to an excessive thickening of the mucous as a consequence of chronic exposure. The acute pulmonary response is typical of a irritant effect with bronchial restriction resulting in increased flow resistance. The chronic effect is similar to that of chronic bronchitis without the involvement of a bacterial infection. These effects have been well reviewed by Costa and Amdur (1996).

88.4  Veterinary effects of sulfur Probably the major health concern of sulfur for ruminants is the association between excessive sulfur ingestion and polioencephalomalacia (PEM), also known as cerebrocortical necrosis. This was first recognized as a disease in sheep and cattle over 40 years ago (Jensen et al., 1956; Terlecki and Markson, 1961). It involves a softening of the gray matter of the brain and is a major disease worldwide (Olkowski, 1997). Clinical signs can occur from a few hours to several weeks after exposure to excessive sulfur. Signs usually include, initially, mild excitation and restlessness accompanied by loss of appetite. Affected animals avoid light and signs may progress to headpressing, rigidity, blindness, violent convulsions, coma, and death. Young animals are particularly badly affected. Removal of affected animals from the source of sulfur generally reduces the severity of clinical signs. In the past, the causes of PEM have been ascribed to a variety of agents, including a lack of vitamin B1 (thiamine) and, more recently, to an excess of dietary sulfur. It has been suggested that the increase in reported cases of PEM over recent years is a result of industrial pollution. However, with the movement away from coal to oil and natural gas, this seems unlikely. For example, according to Beauchamp et al. (1984), the ratio of H2S content of coal, oil, and natural gas, per unit weight of fuel burned, is approximately 35:8:1. Assuming that H2S liberation is a reasonable marker for possible industrial sulfur exposure, this suggests a reduction, rather than an increase, in environmental exposure to sulfur from industrial sources over the years. Several reports have described PEM arising from feeding sheep and/or cattle on elevated levels of sulfur in the diet (Hill and Ebbett, 1997; Jeffrey et al., 1994; Low et al., 1996), in drinking water (Hamlen et al., 1993), as well as a case of sheep being allowed to forage on a field of alfalfa which had been treated with elemental sulfur (Bulgin et al., 1996). Reports of field cases have been duplicated in laboratory studies implicating excessive sulfur ingestion as the cause of PEM (e.g., Gould et al., 1991; McAllister et al., 1992; Sager et al., 1990). PEM was identified in several sheep and cattle farms in England (Jeffrey et al., 1994). It was associated with the use of ammonium sulfate as a feed additive, in place of ammonium

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bicarbonate, as the usual urinary acidifier. After this was discontinued, there were no further cases of PEM. Necropsy of six calves and two lambs from five of these farms showed lesions in the thalamus and striatum, of great severity, unlike in cases of thiamine deficiency. No lesions were found in PEM-affected animals in the cerebellum, hippocampus, or superior colliculi. Symptoms were similar to those already described and there were deaths on three out of five farms. On a cattle farm in New Zealand, PEM was diagnosed following gross and histopathological examination of the brain of deceased animals (Hill and Ebbett, 1997). The cattle had been feeding for 2 months on hay plus a rationed amount of kale (Brassica oleacea) when they were transferred to a field of kale. Within 2 days, neurological signs were observed that were typical of PEM. Twenty-six of 99 heifers (26%) were symptomatic, of which 12 died (12%) and 14 recovered (14%) after removal of the cattle from this field. Chemical analysis of the kale revealed that it contained 8500 mg/kg of sulfur (0.85%, DM, dry matter), which is double the maximum range of recommended dietary needs of cattle for sulfur (0.4%, DM). Examples of sulfur-induced PEM in sheep include an outbreak on a sheep farm in Scotland, after changing from grazing to a ration of pellets containing 0.43% DM sulfur (Low et al., 1996). Clinical signs appeared 15–32 days after changing to the artificial diet, and the incidences were 16 of 46 (35%) for Swaledale lambs and five of 25 (20%) for Scottish blackface lambs. Clinical signs, which were quite unlike those of vitamin B1 deficiency, included depression, blindness, head-pressing, nystagmus, and dorsiflexion of the neck or opisthotonus. In some animals, the severity was such that the sheep either died or were killed in extremis (4/16 Swaledale; 4/5 Scottish blackface). Histopathological examination of the lambs revealed evidence of PEM in the majority of the animals. The mean intake of pellets during the study period was 880 g/head/day (Swaledale) and 760 g/head/ day (Scottish blackface). Because the lambs had an initial average body weight of 20 kg, this intake converts to a food intake of approximately 40 g/kg/day and a sulfur ingestion of approximately 170 mg/kg/day. Administration of vitamin B1 (by injection) did not reverse the clinical signs, but there were no new cases evident after vitamin B1 was given combined with removal of the lambs from the high-sulfur diet. An outbreak of PEM was also reported in cattle which had been drinking water containing a high concentration of sodium sulfate, 7200 ppm vs a recommended optimal level of 1000 ppm, in Canada (Hamlen et al., 1993). The incidence was 11/110 (10%) and mortality of affected cattle was 4/11 (36%). Clinical signs, which first appeared 3 days after exposure to the well, and histopathology (n  3), were the same as those reported above, with the additional findings of extensive thrombosis and vascular necrosis in midbrain and thalamus. The clinical blood chemistry for affected animals appeared to be normal from the standpoint of copper and vitamin B1 levels and transketolase activity

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(a thiamine-dependent enzyme). The PEM dissipated and no new cases arose after the cattle were moved to a water supply with acceptable levels of sulfur. Unusually, old rather than young animals were affected, but this could have resulted from low exposure to sulfur in calves that were nursing. The level of magnesium in the affected well was also high, 1050 vs. 200 ppm recommended. Another example of sulfur toxicity to livestock is a report (Bulgin et al., 1996) of a flock of sheep grazing on a field of alfalfa stubble which had been sprayed 14–16 h previously with an aqueous suspension of 35% elemental sulfur at 53 lbs./acre, active ingredient. (The restricted entry interval for field workers following sulfur use is 24 h, i.e., appropriate personal protective equipment must be worn to enter a treated field within 24 h of a sulfur application.) Within 2–4 h, the sheep became uncoordinated with 91% prostration, despite the sheep being moved to uncontaminated pasture after 2 h, when the problem became apparent. There was 10% (220/2200) mortality after a week, the majority of these sheep (206) dying of acute effects, within 24 h. Surviving ewes were considered fully recovered at 90 days. Necropsy of sheep that died between 2 and 48 h after the onset of clinical signs revealed a rumen pH of 6.0–6.5, a strong smell of rotten eggs (H2S), but no digestive tract lesions. Necropsy of sheep dying between 5 and 30 days showed PEM, consisting of yellow/tan areas of the cerebral cortex caused by neuronal degeneration and cavitation of cortical gray matter. It should be noted that alfalfa normally is moderately rich in sulfur, having a content of ca. 0.4% DM (Olkowski, 1997), without added extraneous sulfur.

Attempts have been made to study the toxicity of sulfur in laboratory experiments. The appearance of clinical signs of PEM in calves fed on a high-sulfate diet coincided with or immediately followed the first odor of H2S in rumen gas (Sager et al., 1990). PEM was not correlated with copper or thiamine deficiency. These findings were extended by Gould et al. (1991), who fed a high-sodium sulfate diet to calves and noted that H2S accumulation in the stomach was significantly higher in animals with signs of PEM than in asymptomatic calves. Microbes in the rumen readily reduce sulfate to sulfide. The findings were extended to sheep by McAllister et al. (1992). Ten lambs were dosed with sodium hydrogen sulfide (0.94 M) every 20 min, administered directly into the esophagus. Clinical signs of PEM developed within 45 min of first dosing, in all lambs, and PEM was identified in four of nine brains examined histologically. All four animals had visual impairment including blindness, dying at 20–96 h. Two lambs had visual impairment without PEM but both died within 90 min of pulmonary congestion and edema (as seen with acute H2S toxicity in the rat), probably before brain lesions had time to develop.

88.5  Human health effects   of sulfur California illness registry data showed 1071 reported cases possibly, probably, or definitely associated with exposure to sulfur between 1992 and 2006 (Table 88.4) (There were 2379 cases associated with sulfur between 1982 and 2006;

Table 88.4  Summary of Cases of Possible, Probable, and Definite Illness Reported to the California Pesticide Illness Registry Involving Exposure to Sulfur as a Primary Pesticide Between 1982 and 2006 Activity

Mixer/loader Applicator

Illness Category Skin

Eye

8

7

44

17

Flagger Mechanic

1

Eye/skin

331

Manufacturing/formulation Transport/storage

3

8 395

Respiratory/ systemic 25

2.3

50

113

10.6

1

0

1

0.1

3

5

9

0.8

1

1

0.1

316

779

72.7

1

1

2

0.2

11

3

17

1.6

5

5

0.5

95 486

119 1071

11.1

2

99

33

Emerg Other

% of total

10

Pack/produce Fieldwork

Total

15 154

1 36

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Table 88.5  Dermatitis complaints following exposure to elemental sulfur Case

Date

County

Narrative

1982–1211

06/29/1982

Madera

Developed skin rash after applying sulfur with wet coveralls.

1983–1252

7/6/1983

Fresno

Tore pants while working with sulfur and got a rash in the area of the tear.

1987–174

02/28/1987

Tulare

A worker complained of a rash after mixing, loading, and applying Kolospray (81% sulfur powder). He had a 2-year history of sensitivity to the material and reported that the rash occurred despite wearing complete safety gear. The treating physician suspected that the dermatitis was due to an allergic reaction and recommended avoiding sulfur powder in the future.

1997–1117

7/4/1997

San Joaquin

A mixer/loader stood on a biplane to control the loading of sulfur dust. During the loading, the dust spontaneously caught fire and he suffered first- and second-degree burns to the face and wrists.

1999–1050

8/26/1999

Solano

Sulfur dust caught fire while being loaded into an airplane hopper. The loader standing on the wing at the time got burned before he jumped off. He suffered first- and second-degree burns on the arms, neck, and the right side of the head, and throat irritation.

2000–485

As two employees loaded sulfur dust into an aircraft’s hopper, the sulfur ignited and burned both employees. This employee suffered first- and second-degree burns on the left ear, arms, and hands.

the complete set of cases was not included because the current coding of the activity variable utilized in Table 88.4 began in 1992.): 486 cases (45.4%) involved sulfur as the primary pesticide associated with the reported illness, the remaining cases involved mixed exposures to sulfur and other pesticides. Of the 1071 total cases, 395 involved cases of dermatitis (36.9%), 154 cases of eye irritation (14.4%), 36 cases of mixed eye/skin irritation (3.4%), and 486 cases involving respiratory or systemic effects (45.4%). By work activity, most noteworthy were the 148 cases (13.8%) involving mixer/load/applicators (pesticide handlers) and the 779 (72.7%) involving field workers.

88.5.1  Occupational exposure To evaluate the typical effects of direct exposure, 198 cases involving handlers (mixer, loader, and applicator cases) reported between 1982 and 2006 are discussed in more detail below. (There were 67 cases reported between 1992 and 2006. These cases constituted 16% of the cases for which sulfur was identified as the primary cause of the reported illness.) Reactions to elemental sulfur can be broken down by illness type.

88.5.1.1  Skin Ninety-one (46%) of the 198 handler cases in 1982–2006 involved dermatitis, with 68 (34.3%) involving isolated dermatitis. The remaining cases involved mixed eye and skin irritation, or mixed respiratory, systemic, and skin

symptoms. Noteworthy cases are discussed below, involving direct exposure, ignition of sulfur during loading or application, or suspected allergy. As illustrated above, some cases appear to have been related to chemical burns or irritation (1982–1211, 1983– 1252, 1997–1117, 1999–1050, 2000–485), while others conceivably could have been related to sulfur allergy (1987– 174). However, provocation tests to confirm the suspected allergy were not documented in any of the cases reported to the California registry. Figures 88.1 and 88.2 show examples of dermatitis of the arm and torso of an exposed worker. These skin lesions appeared within a few minutes of spending 45 min dosing a rose bed with a mixture of elemental sulfur and malathion; they are typical of sulfur. The worker was wearing a shortsleeved shirt without gloves and was sweating profusely in the 95°F heat. This individual is unlikely to be allergic to sulfur since he has applied elemental sulfur on many occasions since this incident (wearing appropriate protective clothing) without experiencing any ill effects. 88.5.1.1.1  Public Domain Literature on Skin Effects of Sulfur The many cases of dermatitis associated with the agricultural use of elemental sulfur suggest that sulfur is a potent skin irritant in humans. Isolated cases have also been reported in applicators and field workers in Washington State. However, standard (epicutaneous) skin irritation tests in laboratory animals for most agricultural formulations have not shown irritation (Matsushita et al., 1977; Table 88.2). In nonstandard tests, however, using subcutaneous injection in the Wistar

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Figure 88.2  A 1 reaction to sulfur – Subject 43 in a California nursery study. A total of five positive reactions to sulfur among 43 subjects. Reprinted with permission from M. A. O’Malley (1997), State of the Art Reviews in Occupational Medicine 12, 327–345.

Figure 88.1  Apparent irritant reaction after a sweaty forearm was contaminated with a mixture of sulfur and malathion. Reprinted with permission from M. A. O’Malley (1997), State of the Art Reviews in Occupational Medicine 12, 327–345.

rat, a 25% aqueous solution of wet-table powder or a 22% solution of lime sulfur caused a level 4 or 5 irritant reaction (1  no reaction; 2  slight hyperemia; 3  hyperemia; 4  marked hyperemia and edema; 5  necrosis.). Using a similar maximization test with the guinea pig, a 1% or 5% aqueous solution of elemental or lime sulfur was a moderately strong allergen (Matsushita et al., 1977). Limited case reports also implicate elemental sulfur as a human contact allergen. Schneider (1978) reported two cases of contact allergy in patients who used medications containing elemental sulfur to treat superficial fungal dermatoses. Both patients had positive patch test reactions to 5% elemental sulfur in various vehicles. A control series was not reported. Wilkinson (1975) reported the case of a professional gardener with a previous history of atopic eczema who developed an eczematous eruption involving the elbow flexures and the right hand. He had a positive patch test reaction to 5% sulfur in petrolatum, but a control series

was not reported. Gregorczyk and Swieboda (1968) described 15 cases of desquamative dermatitis among 425 Polish sulfur miners in which irritant dermatitis due to elemental sulfur may have played a part. Several instances of apparent allergic reaction to 1% elemental sulfur were also observed in a recent study of California nursery workers (O’Malley and Rodriguez, 1998). Allergic contact dermatitis was identified in a hospital investigation of patients suffering from eczematous dermatitis (Vena et al., 1994). Patients were subjected to patch tests with (sodium or potassium) metabisulfite ( S2 O25 – ), − 2– bisulfite ( HSO3 ), or sulfite ( SO3 ). Fifty cases of allergic reaction out of 2894 patients of either sex (1.7%) were reported after exposure to metabisulfite, with 100% crossreactivity between the sodium and potassium salts and with the bisulfite. Only two (4%) of these gave a positive reaction to sulfite. Because metabisulfite is readily converted to bisulfite under aqueous conditions, it is not surprising that they showed cross-reactivity. However, because of the low cross-reactivity toward sulfite, it appears unlikely that sulfite is the ultimate allergen, in vivo, although it is readily formed from the metabisulfite or bisulfite under acidic conditions. It remains possible that some cases of dermatitis resulting from elemental sulfur are due to the subsequent conversion to one of these derivatives. Cases of dermatitis associated with sulfur mineral springs were reported from Taiwan (Sun and Sue, 1995). Over a 10-year period, 44 cases of dermatitis were recorded in visitors to a particular hot springs resort. Two springs were considered, a green sulfur spring (GSS) and a white sulfur one (WSS), and it transpired that all the dermatitis cases had visited the GSS. Of these 44 cases, 32 (70%) had visited the GSS only once, for 10–20 min; 25 (57%)

Chapter | 88  A Toxicological Assessment of Sulfur as a Pesticide

had also visited WSS, without signs of dermatitis; and 24 (55%) had a history of skin diseases prior to visiting the GSS. The chemical and physical properties of the GSS and WSS springs were compared with tap water and microbial infection was ruled out as a cause, since no cultures could be grown from water or affected skin. The principal causes of the dermatitis were considered, by the authors, to be soluble sulfur, which was present at 600, 100, and 80 ppm, in the three water sources, respectively; acid irritation, since pH was 1–2, 4, and 7, respectively, was considered a probable contributory factor. It was also noted that there was a large variation in chloride levels, 3000, 20, and 20 ppm. Other factors that could have contributed to the dermatitis in this hot spring were high temperature (100, 50, and 20°C, respectively) and ammonia nitrogen (200, 0.2, and 0.0, respectively). These disparate pieces of information suggest that active irritants (e.g., sulfuric acid or hydrogen sulfide) or allergens (e.g., sulfites or hydrogen sulfide) may be produced by oxidation or reduction of sulfur. Thus, sulfur may be the precursor of dermal irritants and allergens rather than being one per se.

88.5.1.2  Ocular Irritation Eighty-six (43.4%) of the handler cases involved eye irritation, with 69 cases (34.8%) involving isolated eye irritation.

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Nine of the remaining 17 cases involved eye and skin irritation. Some of the more severe cases are discussed below. As indicated by the above cases, the relationship between the sulfur exposure and the subsequent ocular reaction is usually simple to evaluate, because the irritant or allergic response corresponds directly to the site of contact.

88.5.1.3  Respiratory Tract Illness Classification of respiratory symptoms as systemic or purely local to the respiratory tract is sometimes difficult. Twenty-five (12.6%) of the handler cases involved respiratory illness. Noteworthy cases below included cases of asthma, possible allergic reactions, and exposures to burning sulfur.

88.5.1.4  Systemic Illness Thirty-four of the handler cases (17.1%) involved systemic illness. These included 19 cases (9.6% of the total) without respiratory, eye, or skin effects. Complaints included nonspecific symptoms such as vomiting or feeling shaky (1982–1436), nausea and dizziness (1987–2122 and 1990–1214). Two of the cases also had respiratory symptoms (1987–2122, 1990–1214).

Table 88.6  Eye complaints following contact with sulfur Case

Date

County

Narrative

1987–310

04/01/1987

Sutter

A peach orchard sprayer had exposure to sulfur in eyes and was diagnosed as having chemical keratitis (superficial corneal injury).

1997–824

5/9/1997

San Joaquin

Upon lifting the respirator off his face, sulfur dust got in his left eye. He developed pain and a corneal ulcer.

1995–878

5/4/1995

Napa

A worker was loading sulfur dust into a tractor when the bag suddenly ripped open. Dust got inside his goggles through the air vents and he developed burning, itching, red eyes, and swollen eyelids.

1987–772

4/24/1987

Imperial

Sulfur smoke and fumes were released into the cockpit when the plane caught fire; the cornea of both eyes were burned.

Kern

While dusting grapes with sulfur, he began having an allergic reaction (watery eyes and sneezing) to the sulfur. He apparently has a reaction any time he works around grapes.

1987–1042

Table 88.7  Respiratory illness associated with exposure to sulfur Case

Date

County

Narrative

1984–726

04/25/1984

Kern

Experienced breathing difficulties after working with sulfur, triggered asthma attack.

1985–1134

06/03/1985

Tulare

Loading plane, experienced breathlessness and burning in eyes. He was considered to have a possible allergic reaction to sulfur. (Continued )

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TABLE 88.7  (Continued) Case

Date

County

Narrative

1985–1461

6/24/1985

San Joaquin

A pilot was exposed to burning sulfur when his plane crashed

1986–916

6/5/1986

Stanis-laus A worker developed an asthmatic response when exposed to sulfur dust despite reportedly wearing a face mask respirator.

1987–1042

06/08/1987

Kern

While dusting grapes with sulfur, he began having watery eyes and sneezing, considered an apparent allergic reaction.

1989–1250

05/25/1989

Napa

Worker was applying sulfur with a power sprayer attached to the rear of a tractor. Although he reported wearing a respirator, he developed tightness in the chest and a cough.

1990–1220

06/08/1990

Kern

A pilot applying sulfur to sugar beets crashed when his plane suffered engine failure. The sulfur ignited on impact causing him to inhale the fumes as well as causing burns. He suffered chest tightness, breathing difficulties, and second-degree burns over 20% of his body.

1993–1220

7/22/1993

Solano

A worker was loading sulfur dust from a hopper into an aircraft. The sulfur dust in the hopper shifted causing sulfur dust to be dumped onto the airplane fuselage. The sulfur caught fire from the heat of the engine, burning the worker’s exposed skin. He suffered scattered wheezes, first- and second-degree burns, and singed hair to the face, right arm, and hand.

1994–843

5/21/1994

Napa

A worker developed respiratory problems while dusting sulfur on grapevines. He developed severe shortness of breath and coughing, and was hospitalized for 1.5 days. He has a history of asthma, which the sulfur dust may have aggravated.

2000–101

2/2/2000

Imperial

A crop dusting plane lost power and crashed, igniting its load of sulfur. The pilot was hospitalized for trauma and also hypoxemia, some rhonchi, wheezing, and minor skin burns. He was a smoker, but had no history of asthma or bronchitis.

Table 88.8  Systemic symptoms following sulfur exposure Case

Date

County

Narrative

1982–1436

07/13/1982

Glenn

A nursery worker was exposed to sulfur drifting from an application to an adjacent field. She began to vomit and feel shaky.

1987–2122

5/14/1987

San Mateo

A worker treating ornamentals became ill from sulfur fumes after spilling dust on the muffler of his motorized hand application equipment. Symptoms included nausea, dizziness, vomiting, and wheezing.

1990–1214

05/26/1990

Kern

A worker dusting grapes experienced dizziness, shortness of breath, and a fainting spell.

Table 88.9  Selected cases of injury or illness associated with sulfur combustion Case

Date

County

Narrative Post-crash combustion

1988–3029

09/30/1988

Solano

Pilot had just finished applying one load of sulfur to sugar beets when he crashed and died.

Spontaneous combustion 1994–818

06/11/1994

Kern

Pilot was applying sulfur dust to sugar beets when spontaneous combustion of sulfur caused the cockpit to fill with smoke. While trying to land the plane, it flipped over. He crawled out of the cockpit and was taken to a hospital.

1994–1278

06/25/1994

Sutter

While standing on the wing of a crop-dusting plane, a worker was loading dusting sulfur into the plane’s hopper when the sulfur dust ignited. The resulting explosion knocked him off the plane. He suffered respiratory symptoms and skin burns.

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TABLE 88.9 ������������ (Continued) Case

Date

County

Narrative

1993–1220

07/22/1993

Solano

Worker was loading sulfur dust from a hopper into an aircraft. The sulfur dust in the hopper shifted, causing it to be dumped onto the airplane fuselage. The sulfur caught fire from the heat of the engine and burned the worker’s exposed skin.

1991–1252

07/22/1991

Sonoma

An applicator inhaled smoke from a sulfur/grease fire when a bearing on the application equipment burned. He put out the fire with an extinguisher. The fire did not spread to the hopper full of sulfur.

87–772

04/24/1987

Imperial

Aircraft pilot’s eyes were burned when the plane caught on fire. Sulfur smoke and fumes were released into the cockpit. The cornea of both eyes were burned.

87–2122

05/14/1987

San Mateo

Worker became ill from burning sulfur fumes while dusting sulfur on ornamentals. He apparently spilled dusting sulfur on the muffler of the motorized hand duster during mixing and loading. Symptoms: nausea, dizziness, vomiting, and wheezing.

88.5.1.5  Trauma or Illness Due to Combustion of Sulfur The tendency of sulfur to oxidize makes it prone to combustion. Sulfur combustion was involved in several significant illness episodes reported to the California illness registry. Combustion episodes discussed above resulted in cases of dermatitis (1997–1117, 1999–1050, 2000–485) respiratory illness (1985–1461, 1990–1220, 1993–1220, 2000–10190–1220), ocular irritation (1987–772), and systemic illness (1987–2122). Additional episodes discussed above include both spontaneous combustion and combustion following aircraft accidents (Table 88.9).

88.6  Discussion Environmental exposure to sulfur arises from two main anthropogenic sources, industrial automobile emissions and the use of sulfur in agriculture. This chapter has concentrated on the latter. The extensive use of elemental sulfur as a fungicide in agriculture has, on occasion, led to veterinary problems in animals ingesting toxic levels of sulfur. In addition, sulfur dioxide is used as a fumigant, generally as a preservative for food and drink. In ruminants, sulfur is converted by microorganisms in the rumen to hydrogen sulfide, which is readily absorbed. Sulfide can then inhibit a variety of enzymes involved in oxidative metabolism. It also inhibits respiration by blocking the carotid body and by combining with hemoglobin to produce sulfhemoglobin, thus reducing the oxygen-carrying capacity of the blood. High concentrations of sulfur may lead to secondary thiamine deficits. Sulfite ion is a strong nucleophile and readily binds to thiamine (to the positively charged nitrogen in the thiazole ring), leading to secondary deficits of vitamin B1. Veterinary problems associated with the ingestion of excessive sulfur include polioencephalomalacia (PEM), a severe brain

disease of ruminants. A combination of anecdotal reports of field incidents and laboratory studies has clearly shown that dietary sulfur in these animals needs to be carefully regulated. Elemental sulfur appears relatively inert in both the Buehler and Draize skin test models (see above) but is reported to show moderate capacity to cause sensitization in the guinea pig maximization test (Gregorczyk and Swieboda, 1968). This ambiguous response to sulfur in animal tests is contradicted by use experience, where dermatotoxicity in humans appears to be common. A possible explanation may lie in the transformation of sulfur through oxidation and reduction (Matsushita et al., 1977), which may not readily occur in epicutaneous tests in rodents, which thermoregulate by increased respiration rather than by perspiration. The tendency of sulfur to spontaneously transform under field conditions is underscored by the cases of combustion associated with its use, principally during the summer months in California, where daytime temperatures in the hot, dry inland valleys commonly exceed 100°F (38°C). Current labeling and California regulations prohibit the aerial application of sulfur dust when ambient temperatures exceed 90°F (32°C), in order to reduce combustion incidents due to sulfur. Data from the California Illness Registry do not clearly indicate how many symptoms of sulfur exposure are due to irritant reactions, how many are possibly due to allergic mechanisms, and how many are due to unknown physiologic mechanisms. No cases were recorded to have provocation testing, the only means of clearly distinguishing between irritant and allergic reactions. Five instances of apparent allergic reaction to 1% elemental sulfur, ranging from 1 to 2 /3 [Reactions were scored as 1 (weak reaction, macular erythema), 2 (strong reaction, edematous or vesicular), or 3 (extreme reaction, spreading, bulbous, ulcerative). Equivocal reactions are designated as / (Adams, 1990).], were also observed in a study of California nursery workers. Reactions to sulfur were

1900

correlated with a history of working as a pesticide applicator in the nursery business, but none of the participants had a detailed memory of pesticides they had handled or specifically remembered spraying elemental sulfur (O’Malley and Rodriguez, 1998). Two case reports implicate elemental sulfur as a human contact allergen. Schneider (1978) reported two cases of contact allergy in patients who used medications containing elemental sulfur to treat superficial fungal dermatoses. Both patients had positive patch test reactions to 5% elemental sulfur in series of vehicles, but a control series was not reported. Wilkinson (1975) reported that a professional gardener with a previous history of atopic eczema developed an eczematous eruption involving the elbow flexures and the right hand. He had a positive patch test reaction to 5% sulfur in petrolatum, but a control series was not reported. Gregorczyk and Swieboda (1968) described 15 cases of desquamative dermatitis among 425 Polish sulfur miners in which irritant dermatitis due to elemental sulfur may have played a part.

Conclusion Sulfur continues to be a heavily used fungicide in many parts of agriculture. It’s most notable chemical property is its tendency to spontaneously oxidize. This property is responsible for sulfur’s effects on the eye, skin, and respiratory tract. Safety concerns presented by combustion of sulfur can be mitigated by avoiding its use during periods of high ambient temperature.

Acknowledgments The opinions expressed in this chapter represent the views of the authors and do not necessarily reflect the views and policies of the Department of Pesticide Regulation.

References Adams, R. M. (1990). “Occupational Skin Disease,” Saunders, Philadelphia. Beauchamp, R. O. Jr., Bus, J. S., Popp, J. A., Boreiko, C., and Andjelkovich, D. A. (1984). A critical review of the literature on hydrogen sulfide toxicity. CRC Crit. Rev. Toxicol. 13, 25–97. Bulgin, M. S., Lincoln, S. D., and Mather, G. (1996). Elemental sulfur toxicosis in a flock of sheep. J. Am. Vet. Med. Assoc. 7, 1063–1065. Klaassen, C. D. (ed.) (1996). “Casarett and Doull’s ‘Toxicology, the Basic Science of Poisons’” 5th ed. McGraw–Hill, New York. DPR (1995). “Summary of Pesticide Use Report Data Annual 1993.” Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, CA. DPR (1996a). “Summary of Pesticide Use Report Data Annual 1994.” Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, CA.

Hayes’ Handbook of Pesticide Toxicology

DPR (1996b). “Summary of Pesticide Use Report Data Annual 1995.” Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, CA. Farm Chemicals Handbook. (1998). Volume 84 (R. T. Meister, ed.), Meister Publishing Company, USA. Gador, J., and Motowicka-Terelak, T. (1986). Effect of contamination with sulfur on soil properties and crop yields in a lysimeter experiment: II Effect of elemental sulphur application to the soil on the yields and chemical compositions of some crops. Pamietnik Pulawski 88, 25–38. Gibson, D. M., Kennelly, J. J., and Arherra, F. X. (1987). The performance and thiamine status of pigs fed sulfur dioxide treated highmoisture barley. Can. J. Anim. Sci. 67, 841–854. Gould, D., McAllister, M. M., Savage, J. C. et al. (1991). High sulfide concentration in rumen fluid associated with nutritionally induced polioencephalomalacia in calves. Am. J. Vet. Res. 52, 1164–1169. Gregorczyk, L., and Swieboda, K. (1968). Uber den einfluB von schwefelverbindungen auf die haut und auf die schleimhaute [On the influence of sulfur binding on the skin and mucous membranes]. Polskie Tygognik Lekarski 23, 463–466. Hamlen, H., Clark, E., and Janzen, E. (1993). Polioencephalomalacia in cattle consuming water with elevated sodium sulfate levels: A herd investigation. Can. Vet. J. 34, 153–158. Haynes, R. J., and Swift, R. S. (1986). Effects of soil amendments and sawdust mulching on growth, yield and leaf nutrient content of highbush blueberry [Vaccinium corymbosum cultivar Bluecrop] plants. Scientia Hortic 29(3), 229–238. Hill, F. I., and Ebbett, P. C. (1997). Polioencephalomalacia in cattle in New Zealand fed chou moellier (Brassica oleracea). New Zealand Vet. J. 45, 37–39. Jeffrey, M., Duff, J. P., Higgins, R. J., Simpson, V. R., Jackman, R., Jones, T. O., Mechie, S. C., and Livesey, C. T. (1994). Polioencephalomalacia associated with the ingestion of ammonium sulphate by sheep and cattle. Vet. Rec. 134, 343–348. Jensen, R., Griner, L. A., and Adams, O. R. (1956). Polioencephalomalacia of cattle and sheep. J. Am. Vet. Med. Assoc. 129, 311–321. Low, J. C., Scott, P. R., Howie, F., Lewis, M., Fitzsimons, J., and Spence, J. A. (1996). Sulfur-induced polioencephalomalacia in lambs. Vet. Rec. 138, 327–329. Majule, A. E., Topper, C. P., and Nortcliff, S. (1997). The environmental effects of dusting cashew (Anacardium occidentale L.) trees with sulphur in southern Tanzania. Trop. Agric. 74(1), 25–33. Matsushita, T., Yoshioka, M., Aoyama, K., Arimatsu, Y., and Nomura, S. (1977). Experimental study on contact dermatitis caused by fungicides benomyl and thiophanate-methyl. Ind. Hlth. 15(3–4), 141–148. McAllister, M. M., Gould, D. H., and Hamar, D. W. (1992). Sulfideinduced polioencephalomalacia in lambs. J. Comp. Pathol. 106, 267–278. Olkowski, A. A. (1997). Neurotoxicity and secondary metabolic problems associated with low to moderate levels of exposure to excess dietary sulfur in ruminants: A review. Vet. Human Toxicol. 39, 355–360. O’Malley, M. A., and Rodriguez, H. -P. (1998). “Contact Dermatitis in California Nursery Workers: Part II. Pilot Field Study.” California EPA, DPR, Worker Health and Safety Branch HS-1767. Sager, R. L., Hamar, D. W., and Gould, D. (1990). Clinical and biochemical alterations in calves with nutritionally induced polioencephalomalacia. Am. J. Vet. Res. 51, 1969–1974. Schneider, H. G. (1978). Schwefelallergie [Sulfur allergy]. Hautarzt 29, 340–342.

Chapter | 88  A Toxicological Assessment of Sulfur as a Pesticide

Shapiro, R. (1977). Genetic effects of bisulfite (sulfur dioxide). Mutat. Res 39, 149–176. Studdert, V. P., and Labuc, R. H. (1991). Thiamine deficiency in cats and dogs associated with feeding meat preserved with sulphur dioxide. Aust. Vet. J. 68, 54–57. Sun, C. C., and Sue, M. S. (1995). Sulfur spring dermatitis. Contact Dermat 32, 31–34.

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Terlecki, S., and Markson, L. M. (1961). Cerebrocortical necrosis in cattle and sheep. Vet. Rec. 73, 2327. Vena, G. A., Foti, C., and Angelini, G. (1994). Sulfite contact allergy. Contact Dermat 31, 172–175. Wilkinson, D. S. (1975). Sulfur sensitivity. Contact Dermat 1, 58.

Chapter 89

Cyprodinil: A Fungicide of the Anilinopyrimidine Class Felix Waechter, Edgar Weber, Thomas Hertner, and Ursula May-Hertl Syngenta Crop Protection AG, Basle, Switzerland

89.1  Introduction Anilinopyrimidines are a new chemical class of fungicides that are highly active against a broad range of fungi. Three compounds are currently on the market: Cyprodinil (Syngenta Crop Protection AG), Mepanipyrim (Kumiai Chemical Industry Co., Ltd.), and Pyrimethanil (AgrEvo GmbH). They all have similar structures and differ only with respect to their substituent at position 4 of the pyrimidine ring, which is a cyclopropyl group for Cyprodinil, a 1-propynyl group for Mepanipyrim, and a methyl group for Pyrimethanil (Figure 89.1). The biological mode of action includes inhibition of methionine biosynthesis and secretion of hydrolytic enzymes. Anilinopyrimidines show no cross-resistance with other fungicide groups. Extensive field monitoring has shown reliable performance of anilinopyrimidine products over many years and no cases of practical resistance have been detected under commercial conditions in all key target pathogens including Botrytis and Venturia. However, under conditions of forced selection, a potential resistance risk of these pathogens could be demonstrated in field and laboratory trials. This chapter addresses the toxicological profile and human safety aspects of Cyprodinil as a representative example of anilinopyrimidines. Where published data are available, reference is made to the toxicological profile of the other two marketed anilinopyrimidines, Mepanipyrim and Pyrimethanil.

4-cyclopropyl-6-methyl-N-phenyl-2-pyrimidinamine (ac­c­o­rding to CA).

89.2.2  Synonyms The common name Cyprodinil (ISO) is in general use. Trade names include Unix (Syngenta), Chorus (Syngenta), Stereo (Syngenta), Switch (Syngenta), and Vangard (Syngenta USA). The CAS registry number is 121552–61–2 and the developmental code is CGA 219417.

N

NH N

Cyprodinil

N

NH N

Mepanipyrim

N

NH

89.2  Identity, Properties, and Uses 89.2.1  Chemical Name Cyprodinil (Figure 89.1) is (4-cyclopropyl-6-methylpyrimidin-2-yl)-phenyl-amine (according to IUPAC) or Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

N

Pyrimethanil Figure 89.1  Chemical structures of Cyprodinil, Mepanipyrim, and Pyrimethanil.

1903

1904

89.2.3  Physical and Chemical Properties Cyprodinil has the molecular formula C14H15N3 and a molecular weight of 225.3. It is a white crystalline solid with a melting point of 75.9°C. The vapor pressure of Cyprodinil at 25°C is 5.1  104 Pa (crystal modification A) or 4.7  104 Pa (crystal modification B). Its solubility in water at 25°C is 20 mg/l (pH 5.0), 13 mg/l (pH 7.0), or 15 mg/l (pH 9.0). It is quite soluble in most organic solvents. The partition coefficient log POW (n-octanol/ water) is 3.9 at pH 5.0 or 4.0 at pH 7.0 and pH 9.0 (Heye et al., 1994).

89.2.4  History, Formulations, and Uses Cyprodinil was introduced in France in 1993 for application on cereal grains. It is currently used as a foliar fungicide in cereal grains, grapes, pome fruit, stone fruit, strawberries, vegetables, field crops, and ornamentals, and as a seed dressing on barley. Among others, products that contain Cyprodinil are marketed in most European countries, in North America, and in Japan. Cyprodinil controls a wide range of pathogens such as Pseudocercosporella herpotrichoides, Erysiphe spp., Pyrenophora teres, Rhynchosporium secalis, Septoria nodorum, Botrytis spp., Alternaria spp., Venturia spp., and Monilinia spp. Formulations for foliar applications are of the WG (water dispersible molecules) or EC (emulsifiable concentrates) type; those for seed dressing are of the FS (flowable concentrate for seed treatment) type. Important mixing partners are triazoles, fludioxonil, fenpropidin, and acibenzolar-S-methyl.

89.3  Biological mode of action The anilinopyrimidine fungicides Cyprodinil, Pyrimethanil, and Mepanipyrim were identified as inhibitors of methionine biosynthesis in liquid cultures of Botrytis cinerea. In filamentous fungi, such as Neurospora grassa and Aspergillus nidulans, the cystathionine pathway has been established as the major route of homocysteine and methionine biosynthesis (Yamagata, 1989). Within this pathway, cystathionine -lyase, which catalyzes the synthesis of homocysteine from cystathionine, was identified as a target for these anilinopyrimidines (Fritz et al., 1997; Masner et al., 1994). The plant cell-wall degrading enzyme-secretion system of pathogenic fungi was identified as a second target for anilinopyrimidine fungicides. Cell-wall digesting enzymes of pathogenic fungi are thought to play a critical role in the early step of plant infection, that is, in the absence of secretion, penetration of plant tissue by fungal mycelium is inhibited (Milling and Richardson, 1995; Miura et al., 1994).

Hayes’ Handbook of Pesticide Toxicology

89.4  Absorption, distribution, metabolism, and excretion 89.4.1  Toxicokinetics in Rats 89.4.1.1  Oral Administration The distribution and depletion kinetics in male and female rats were investigated with [U-14C]phenyl-cyprodinil and [U-14C]pyrimidyl-cyprodinil. The labeled test substance was administered to male and female rats at single oral doses of 0.5 (only phenyl-labeled compound) and 100 mg/kg. Additional animals of both sexes received 14 consecutive doses of the nonlabeled test article followed by a single treatment with the phenyl-labeled compound at 0.5 mg/kg. The biliary excretion was investigated in bile-fistulated males treated with 100 mg/kg of the phenyllabeled test article. Independent of dose level and sex, Cyprodinil was rapidly absorbed from the gastrointestinal tract into systemic circulation. The excretion pattern was essentially independent of the sex, dose level, pretreatment, or label of the test substance administered. Approximately 52–68% of the administered dose was excreted in the urine, whereas 33–45% was found in the feces. Total excretion reached 92–97% of the administered dose within 48 h. In bilefistulated rats, 39, 35, and 14% of the radioactivity were excreted via bile, urine, and feces, respectively. A comparison of these values with those obtained with the nonfistulated animals suggests that a significant amount eliminated via the feces was absorbed and reentered the intestinal tract by biliary excretion. A smaller amount seemed to be reabsorbed from the intestine and eliminated via the kidneys. Residues in tissues were generally low and rapidly depleted with no evidence for an accumulation or retention of radioactivity.

89.4.1.2  Dermal Administration The dermal absorption of Cyprodinil formulated as a WGtype product was determined in rats. The compound was applied for 6 h at two dose levels: the lower dose represented a typical spray dilution; the higher dose represented the highest applicable concentration. Residues in blood, skin, urine, and feces were determined over 48 h. Blood levels reached a maximum within 1–2 h of exposure and decreased rapidly. The total amount absorbed within 48 h was roughly 20% of the administered low dose and about 3% of the high dose.

89.4.2  Metabolic Pathways in Rats The metabolic fate of Cyprodinil was investigated in urine, feces, and bile samples obtained from the toxicokinetic investigations. In urine and feces, there were no differences

Chapter | 89  Cyprodinil: A Fungicide of the Anilinopyrimidine Class

in the metabolite patterns of the phenyl and pyrimidyl labels, that is, there were no metabolites identified in these excreta that indicated a cleavage of the C–N–C bridge between the phenyl and pyrimidyl ring. Qualitatively, the metabolic pathways were not influenced by the dose level, by pretreatment with unlabeled Cyprodinil, or by the sex of the animals. Seven urinary, eight biliary, and three fecal metabolites were identified, which in total accounted for 65–80% of the administered radioactivity. Cyprodinil was almost completely metabolized. No unchanged parent molecule could be found in urine, whereas minor amounts of unchanged Cyprodinil were found in feces. Most of the administered Cyprodinil was metabolized by sequential oxidation of the phenyl and pyrimidine ring (Figure 89.2). The major phase 1 metabol­ ite was identified as 4-cyclopropyl-5-hydroxy-6-methyl-N(4-hydroxy)-phenyl-2-pyrimidinamine (metabolite 2). This metabolite was excreted in the urine as -glucuronic acid conjugate as well as mono- and disulfuric acid conjugates. Although female rats formed the monosulfate almost exclusively, the males excreted equal amounts of the mono- and disulfate. Further oxidation of the methyl group led to the formation of 4-cyclopropyl-5-hydroxy-6-hydroxymethylN-(4-hydroxy)-phenyl-2-pyrimidinamine (metabolite 3), which was excreted in the urine in unconjugated form. Alternative pathways proceeded either by sequential oxidation of the phenyl ring to the 4-hydroxy and 3,4dihydroxy derivatives (metabolites 4 and 5), followed by oxidation of the methyl group (metabolite 6), or started with NH

NH

NH

NH2

HO

Cyprodinil

NH

NH

N OH

Metabolite 1

NH

Metabolite 8

NH

N

N

Metabolite 9

N

N

OH

Metabolite 4

HO

N

N

HO

NH

HO OH

N

OH

Metabolite 3

89.4.4  Toxicokinetics and Metabolism in Laying Hens

N

HO

OH

Metabolite 6

Figure 89.2  Phase 1 metabolism of Cyprodinil in the rat.

89.4.3  Toxicokinetics and Metabolism in Lactating Goats

OH

Metabolite 5

N

N

NH

HO OH

N

N

HO

Metabolite 2

NH

N

H2N

hydroxylation of the methyl group as the first oxidation step (metabolite 9). Urinary and biliary metabolites were found to be conjugated with -glucuronic acid and sulfuric acid. The major metabolites identified in feces were the 5hydroxypyrimidine derivative of Cyprodinil (metabolite 1) and metabolite 4. Metabolites 1 and 4 also were present in conjugated form in urine and bile. In addition, liver and kidney tissue residues were analyzed for metabolites 12 h after single oral administration of [2-14C]pyrimidyl-Cyprodinil at 100 mg/kg to male rats. During this period, 17.0% of the total applied dose was eliminated via urine. The identified liver and kidney metabolites essentially confirmed the metabolic pathways proposed upon analysis of metabolites isolated from excreta. However, two additional metabolites were found in liver and/or kidney tissue, but not in excreta. Metabolite 7 was identified as ring-hydroxylated N-phenyl-guanidine, a breakdown product of the pyrimidine ring moiety. Metabolite 8 (i.e., 4-cyclopropyl-6-methyl-pyrimidine-2ylamine) demonstrated a cleavage of the parent molecule between the pyrimidine and the phenyl ring. Metabolite 7 was found exclusively in the liver, where it represented the major metabolite. Minor amounts of metabolite 8 were found in the liver and kidneys.

[U-14C]phenyl or [2-14C]pyrimidyl labeled Cyprodinil were orally administered in gelatin capsules once daily to lactating goats for 4 consecutive days at dose levels of about 0.2 and 10 mg/kg. The absorption and excretion of radioactivity were measured over 78 h. Six hours after the last dose, the animals were sacrificed and the tissue residues were determined. Independent of the label and the dose administered, Cyprodinil was absorbed in goats to a lesser extent and more slowly than in rats. The major route of excretion was in urine and feces, whereas excretion via the milk was minimal. Residues of radioactivity in edible tissues were generally low. The metabolic pathways of Cyprodinil in lactating goats were similar to those observed in the rat.

N

Metabolite 7

N

HO

N

N

1905

[U-14C]phenyl or [2-14C]pyrimidyl labeled Cyprodinil were orally administered in gelatin capsules to laying hens once daily for 4 consecutive days at dose levels of about 0.4 and 19 mg/kg. The excretion of radioactivity was measured over 78 h. Six hours after the last dose, the animals were sacrificed and the tissue residues were determined.

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1906

In laying hens, Cyprodinil was rapidly and completely eliminated. Residues of radioactivity in eggs and edible tissues were very low. The distribution and excretion pattern was independent of the dose and the site of labeling. The metabolic pathways of Cyprodinil in laying hens were similar to those observed in the rat.

89.5  Toxicity to laboratory animals 89.5.1  Acute Toxicity The acute toxicity profile of Cyprodinil is summarized in Table 89.1. According to these data, Cyprodinil is unlikely to present an acute hazard in normal use (Class III, according to the WHO hazard classification scheme). Cyprodinil is nonirritant to skin and eye when classified according to the EC Directive 83/467, but it may cause sensitization by skin contact.

89.5.2  Subchronic Toxicity The following subchronic toxicity studies were conducted: l l l l

Rat, 28-day gavage Rat, 28-day dermal Rat, 90-day feeding Mouse, 90-day feeding Dog, 90-day feeding Dog, 12-month feeding

l l

Cyprodinil was administered by gastric intubation at 0, 10, 100, and 1000 mg/kg to groups of 10 male and female

Table 89.1  Acute Toxicity of Cyprodinil Parameter

Species, strain, number, and sex

Result

Acute oral LD50

Rat, Tif RAIF (SPF), 5 m  5 f Mouse, Tif MAG (SPF), 5 m  5 f

Greater than 2000 mg/kg Greater than 5000 mg/kg

Acute dermal LD50

Rat, Tif RAIf (SPF), 5 m  5 f

Greater than 2000 mg/kg

Acute inhalation LC50

Rat, Tif RAIf (SPF), 5 m  5 f

Greater than 1200 mg/m3a

Skin irritation

Rabbit, New Zealand White, 3 f

Nonirritantb

Eye irritation

Rabbit, New Zealand White, 3 m

Nonirritantb

Sensitization (maximization test)

Guinea pig, Pirbright White, 10 m  10 f

Sensitizingb

Maximum attainable concentration. Classified according to EC Directive 83/467.

b

89.5.2.2  Rat, 28-Day Dermal Groups of five male and female Tif: RAIf (SPF) rats were dermally treated with Cyprodinil at doses of 0, 5, 25, 125, and 1000 mg/kg for 6 h per day, 5 days per week for 4 weeks (OECD Guideline 410). Except for some clinical observations of doubtful relationship to treatment (piloerection, dyspnea, hunched posture) and pathological changes at the application site, which were attributed to the application procedure, no effects of treatment were observed. In particular, Cyprodinil did not induce local irritations or effects on clinical chemistry parameters.

89.5.2.3  Rat, 90-Day Feeding

89.5.2.1  Rat, 28-Day Gavage

a

Tif-RAIf (SPF) rats for 5 days per week for 4 weeks. The main findings in both sexes were hepatomegaly, as expressed by increased liver weight at 100 and 1000 mg/ kg with a corresponding hepatocellular hypertrophy, and increased thyroid weights associated with follicular cell hypertrophy at 1000 mg/kg. The findings in the thyroid were considered to be secondary to liver stimulation. Further changes at 1000 mg/kg included decreased body weight development and food consumption, hypochromasia, and increased serum concentrations of albumin, globulin, bilirubin, phospholipids, and cholesterol.

Groups of 10 male and female Tif:RAIf (SPF) rats received Cyprodinil at dietary concentrations of 0, 50, 300, 2000, and 12,000 ppm for 90 days. An additional control and high-dose group was kept for a 4-week recovery period. Treatment resulted in partially reversible reduction of body weight development and food consumption at 12,000 ppm. Slightly elevated serum concentrations of cholesterol and phospholipids occurred in both sexes treated at 2000 and 12,000 ppm, and minimally increased serum activities of hepatic enzymes (alkaline phosphatase, -glutamyltranspeptidase) were noted at 12,000 ppm. These changes were completely reversible after 4 weeks of recovery. Slightly elevated weights were found for the liver in both sexes at 2000 and 12,000 ppm. Relative kidney weights were increased in both sexes in the top dose group and relative thyroid weights were increased in the animals of the top dose group and males of the 2000-ppm dose group. In the recovery group, only the thyroid weights remained slightly elevated. Histopathology revealed treatment-related effects in the kidneys (chronic tubular lesion) as well as in the liver (hepatocellular hypertrophy, foci of necrotic hepatocytes, hepatocellular inclusion bodies), thyroid (hypertrophy of thyroid follicles), and pituitary gland (hypertrophy of pituitary cells). The changes in the thyroid and pituitary gland might be related to liver stimulation. Except for the renal lesions, all changes were at least partially reversible within the 4-week recovery period. The elevated serum concentrations of cholesterol and phospholipids, seen after treatment with Cyprodinil for 90 days

Chapter | 89  Cyprodinil: A Fungicide of the Anilinopyrimidine Class

in rats, are different from those reported by Terada et al. (1998a,b) for the structurally closely related anilinopyrimidine Mepanipyrim. Treatment of male rats at 4000 ppm Mepanipyrim for 3 weeks decreased the serum cholesterol, triglyceride, and phospholipid levels (Terada et al., 1998b). In addition, these authors reported hepatocellular fatty vacuolation after treatment for 13 weeks at 200 ppm and above (Terada et al., 1998a), an effect that did not occur in Cyprodinil-treated rats at doses up to 12,000 ppm.

89.5.2.4  Mouse, 90-Day Feeding Cyprodinil was administered to groups of 10 male and female Tif:MAGf (SPF) mice over 90 days at dietary concentrations of 0, 500, 2000, and 6000 ppm. The animals tolerated the subchronic dietary administration without mortality or overt clinical signs. Body weight gain, food consumption, and hematological profile remained unaffected by the treatment. The liver was the main target organ of tox­icity. Treatment of males at 6000 ppm caused increased absolute and relative liver weight. Histopathological changes at 2000 ppm and above comprised single cell necroses in males and depletion of glycogen in females. Clinical chemistry parameters were not investigated.

89.5.2.5  Dog, 90-Day Feeding Groups of four male and female beagle dogs received Cyprodinil at dietary concentrations of 0, 200, 1500, 7000, and 20,000 ppm for 90 days (OECD Guideline 409). Treatment with Cyprodinil was well tolerated even at very high-dose levels. The animals reacted to the treatment mainly by reduced food consumption and reduced body weight gain. In addition, increased numbers of blood platelets were recorded in the females treated at the top-dose level of 20,000 ppm. The laboratory examinations indicated no deviations of possible toxicological significance. In particular, the treatment did not affect the plasma concentrations of cholesterol. This observation is in line with the lack of changes in clinical chemistry parameters in Mepanipyrim-treated dogs (Terada et al., 1998a). The lipofuscin deposition in Kupffer cells and hepatocytes, as observed with Mepanipyrim (Terada et al., 1998a), was not seen with Cyprodinil.

89.5.2.6  Dog, 12-Month Feeding Groups of four male and female beagle dogs received Cyprodinil at dietary concentrations of 0, 25, 250, 2500, and 15,000 ppm (OECD Guideline 452). The animals tolerated the treatment without clinical signs or mortality and no effects on laboratory parameters were encountered. Reactions to the treatment mainly consisted of reduced body weight gain and food consumption at the top-dose level. In addition, the livers of the male dogs dosed at 15,000 ppm contained minor amounts of lipofuscin-like pigments. Lipofuscinosis also has been reported upon treatment of dogs with Mepanipyrim (Terada et al., 1998a).

1907

89.5.3  Chronic Toxicity and Oncogenicity The following chronic toxicity and oncogenicity studies were conducted: Mouse, 18-month feeding Rat, 24-month feeding

l l

89.5.3.1  Mouse, 18-Month Feeding Cyprodinil was administered to groups of 60 male and female Tif:MAGf (SPF) mice over 18 months at dietary concentrations of 0, 10, 150, 2000, and 5000 ppm (OECD Guideline 451). The animals tolerated the chronic treatment without clinical signs. Most probably due to a reduced body weight development, survival was slightly increased in the top-dose group. Slightly increased liver and kidney weights were noted in the top-dose-group animals. Histopathological examinations revealed increased incidences of slight to moderate hyperplasia of the exocrine pancreas in the top-dosegroup males. The incidence and distribution of neoplastic changes were similar in treated and untreated control groups, and remained within the range of the historical controls.

89.5.3.2  Rat, 24-Month Feeding Cyprodinil was administered to groups of 80 male and female Tif:RAIf (SPF) rats at dietary concentrations of 0, 5, 75, 1000, and 2000 ppm over 24 months (OECD Guideline 453). The chronic dietary administration of Cyprodinil was tolerated without clinical signs or treatment-related mortality. Hematological examinations revealed a slight prolongation of the prothrombin time in the top-dose-group males only during the first half of the treatment period. At the 12-month interim sacrifice, higher kidney-to body-weight ratios were observed for both sexes at 2000 ppm. At terminal sacrifice, males showed increased absolute and relative liver weights at the top-dose level and hepatic sinusoidal cystic dilatation at 1000 ppm and above. The incidence and distribution of neoplastic lesions gave no indication of a carcinogenic effect. In particular, there was no induction of thyroid follicular cell tumors, which were found to be elicited in rats by treatment with the structurally closely related anilinopyrimidine Pyrimethanil (Hurley, 1998). Enhancement of hepatic thyroid hormone metabolism and excretion are considered to be the mode of action of thyroid tumorigenesis (Hurley, 1998).

89.5.4  Effects on Liver Xenobiotic Metabolizing Enzymes in the Rat In subchronic oral toxicity studies in the rat, Cyprodinil caused increased liver weights and hepatocellular hypertrophy. Its possible effects on hepatic xenobiotic metabol­ izing enzymes were investigated in an exploratory 28-day study where male Tif RAIf (SPF) rats were treated by oral intubation at dose levels of 0, 100, and 1000 mg/kg

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1908

per day (Table 89.2). The treatment was without effect on body weight, but caused increased liver weights to 123 and 160% of control at 100 and 1000 mg/kg, respectively. Increased microsomal protein and cytochrome P450 contents at 1000 mg/kg were accompanied by an indu­ction of the investigated cytochrome P450-dependent monooxygenase activities, that is, ethoxyresorufin O-deethylase and pentoxyresorufin O-depentylase as well as lauric acid 11- and 12- hydroxylase. The most prominent effect was a nearly 30-fold induction of pentoxyresorufin O-depentylase, a marker enzyme activity for phenobarbitone-indu­ cible cytochrome P450 isoenzyme CYP2B1 (Whitlock and Denison, 1994). Cytosolic glutathione S-transferase activity toward 2,4-dinitrochlorobenzene was induced two- and 3.5-fold at the lower and higher dose level, respectively, whereas cyanide insensitive fatty acyl CoA  oxidation, a marker for hepatic peroxisome proliferation (Lake and Lewis, 1993), was not affected. According to these data,

Table 89.2  Effects of Cyprodinil on Liver Xenobiotic Metabolizing Enzymes in the Male Rata Parameter

Dose (mg/kg body weight/day) 0

100

1000

Final body weight (g)

341 (10)

357 (11)

330 (26)

Absolute liver weight (g)

10.1 (0.5)

12.5** 16.2*** (0.9) (1.7)

Microsomal protein (mg/g liver)

25.9 (1.3)

25.3 (4.6)

35.0* (5.0)

Cytosolic protein (mg/g liver)

127 (5)

123 (5)

114*** (3)

Cytochrome P450 (nmol/g liver)

20.6 (2.7)

18.1 (3.3)

46.6*** (11.8)

Ethoxyresorufin O-deethylase (nmol/min g1 liver)

4.40 (1.25)

9.40** 43.5*** (3.86) (10.1)

Pentoxyresorufin O-depentylase (nmol/min g1 liver)

1.18 (0.12)

3.56** 33.6*** (1.92) (6.95)

Lauric acid 11-hydroxylase (nmol/min g1 liver)

22.5 (1.8)

22.8 (5.4)

46.0*** (8.8)

Lauric acid 12-hydroxylase (nmol/min g1 liver)

14.9 (2.6)

21.0* (5.2)

45.7*** (7.1)

Glutathione S-transferase (mol/min g1 liver)

87.4 (32.3)

184** (49)

305*** (36)

Fatty acid -oxidation (nmol/min g1 liver)

1034 (204)

1129 (128)

1139 (160)

a

Mean values; standard deviations are given in parentheses. Asterisks indicate statistical significant difference: * p  0.05, ** p  0.01, *** p  0.001. Five animals were used in each dosage group.

the hepatomegaly induced by Cyprodinil in the rat can be interpreted as an adaptation to a functional load.

89.5.5  Effects on Liver and Plasma Lipids   in the Rat Upon subchronic feeding to rats, Cyprodinil caused increased plasma concentrations of cholesterol and phospholipids. In an exploratory study, male Tif:RAIf (SPF) rats were treated with Cyprodinil for 28 days by oral intubation at dose levels of 0, 100, and 1000 mg/kg per day. After a 20-h fasting period, the animals were sacrificed and analyzed for the content of liver free cholesterol and cholesterol esters and for the concentration of serum total cholesterol, as well as for the concentration of cholesterol bound to highdensity lipoprotein (HDL), low-density lipoprotein (LDL), and very-low-density lipoprotein (VLDL) (Table 89.3). In the liver, the concentration of free cholesterol level remained unaffected and the cholesterol ester concentration was, if at all, slightly reduced in both treatment groups. In the rat, cholesterol is mainly transported by HDL and to a lower extent also by LDL, whereas VLDL is of minor importance (Carroll and Feldman, 1989). Serum total chol­ esterol concentration was increased by a factor of about 2 at 1000 mg/kg. As expected, cholesterol concentrations were increased in all three serum lipoprotein fractions of animals treated with 1000 mg/kg, whereby the higher total serum cholesterol level could largely be attributed to increased HDL and LDL levels. These data confirm that, upon subchronic administration to rats, Cyprodinil interacts with lipid homeostasis. In special investigations, the structurally related anilinopyrimidine Mepanipyrim was also shown to interfere with lipid homeostasis in the rat (Terada et al., 1998b). However, the effects elicited by the two anilinopyrimidines are quite different (Table 89.4). Mepanipyrim caused fatty liver that comprised increased liver cholesterol, phospholipid, and triglyceride concentrations. Cyprodinil was without an effect on liver cholesterol concentration (liver phospholipid and triglycerides were not measured) and did not cause fatty liver. In blood, Mepanipyrim decreased cholesterol and high-density lipoprotein cholesterol, phospholipid, and triglyceride concentrations, whereas Cyprodinil caused increased cholesterol concentration, cholesterol concentrations in high-, low-, and very-low-density lipoprotein fractions, and phospholipid concentrations. It was suggested that the fatty liver induced by Mepanipyrim in the rat is the result of an inhibition of intracellular transport of very-low-density lipoproteins from the Golgi apparatus to the cell surface (Terada et al., 1999). This hypothesis is in accordance with its presumed mode of action in pathogenic fungi, where the compound was shown to block intracellular trafficking and secretion of plant cell-wall digesting enzymes (Miura et al., 1994).

Chapter | 89  Cyprodinil: A Fungicide of the Anilinopyrimidine Class

Table 89.3  Effects of Cyprodinil on Liver Free Cholesterol, Liver Cholesterol Esters, and Serum Cholesterol in the Male Rata Parameter

Dose (mg/kg body weight/day)

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Table 89.4  Cyprodinil and Mepanipyrim: Comparison of Effects on Selected Lipid Parameters in Liver and Blood Following Subchronic Treatment of Rats Parameter

Cyprodinila

Mepanipyrimb

Liver free cholesterol

No effect

No effect

Liver cholesterol esters

No effect

↑

0

100

1000

Liver free cholesterol

8.3

7.7

8.2

(g/mg homogenate protein)

(0.5)

(0.5)

(0.5)

Liver total cholesterol

No effect

↑

Liver cholesterol esters

1.8

1.4

1.3

Liver phospholipid

↑

(g/mg homogenate protein)

(0.2)

(0.2)

(0.2)

Not measured

Liver triglyceride

1.78

1.90

3.89***

Not measured

↑

Serum total cholesterol (mol/ml serum)

(0.22)

(0.30)

(0.74)

↑

↓

High-density lipoprotein cholesterol

1.24

1.36

2.23***

Plasma or serum total cholesterolc

↑

↓

(mol/ml serum)

(0.09)

(0.24)

(0.42)

Plasma or serum phospholipidc

Low-density lipoprotein cholesterol

0.64

0.67

1.86***

Serum high-density

↑

↓

(mol/ml serum)

(0.16)

(0.15)

(0.34)

Very-low-density lipoprotein cholesterol

0.020

0.036

0.128

(mol/ml serum)

(0.021) (0.014) (0.110)

Lipoprotein cholesterol a

Cyprodynil was administered to Tif:RAIf (SPF) male rats via gavage for 4 weeks at a rate of 1000 mg/kg per day. b Mepanipyrim was administered to F344/DuCrj male rats via diet for 3 weeks at a rate of 4000 ppm. Data from Terada et al. (1998b). c Cyprodinil: plasma; Mepanipyrim: serum.

a

Mean values; standard deviations are given in parentheses. Asterisks indicate statistical significant difference: * p  0.05, ** p  0.01, *** p  0.001. Five animals were used in each dosage group.

The mechanism by which Cyprodinil induces increased blood cholesterol and phospholipid concentrations in the rat is not known. However, in vitro investigations with Cyprodinil in primary cultured rat hepatocytes showed that the hypercholesterolemia observed with this compound was neither due to increased hepatic cholesterol synthesis nor a consequence of inhibition of bile acid synthesis or bile acid transport (data not shown).

89.5.6  Effects on Reproduction and Development Possible effects of Cyprodinil on reproduction and development were investigated in the following studies: Rat, teratogenicity l Rabbit, teratogenicity l Rat, two-generation reproduction l

89.5.6.1  Rat, Teratogenicity Cyprodinil was orally administered by gastric intubation to groups of 20–22 pregnant Tif:RAIf (SPF) rats from day

6 to day 15 of gestation at daily doses of 0, 20, 200, and 1000 mg/kg per day. The dams were sacrificed on day 21 of gestation and the fetuses were removed, weighed, sexed, examined for external malformations, and subjected to visceral and skeletal examination. In the dams, no treatment-related mortality occurred and no clinical signs of possible relevance were noted. Body weight development and food consumption were significantly reduced in the top-dose animals. No treatment-related changes were noted upon necropsy. No treatment-related effects were observed on pregnancy rate, corpora lutea, implantations, early or late resorptions, sex, or number of fetuses. Weight reduction and reduced ossification of digits and metacarpalia were observed in the top-dose group. These findings were considered to be a consequence of the observed maternal toxicity at top-dose treatment. In conclusion, the results of this study gave no indication of teratogenic effects.

89.5.6.2  Rabbit, Teratogenicity Groups of 17–18 pregnant Russian rabbits were orally treated by gavage from day 7 to day 19 of gestation at daily doses of 0, 5, 30, 150, and 400 mg/kg per day. The does were sacrificed on day 29 of gestation and the fetuses were removed by hysterectomy, weighed, sexed, examined for external malformations, and subjected to visceral and skeletal examination.

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Neither mortality nor treatment-related clinical signs occurred in the does. The body weight development and the food consumption of the top-dose animals was reversibly reduced during the treatment period. No treatment-related changes were noted upon necropsy. No treatment-related effects were observed on pregnancy rate, corpora lutea, implantations, early or late resorptions, sex, or number of fetuses. The fetal weights were similar in treated and untreated animals, and the external, visceral, and skeletal examinations of the fetuses revealed no treatment-related effects. In conclusion, the results of this study gave no indication of teratogenic effects.

89.5.6.3  Rat, Two-Generation Reproduction Cyprodinil was administered to groups of 30 male and female Tif:RAIf (SPF) rats at dietary doses of 0, 10, 100, 1000, and 4000 ppm over two generations. In the F0 generation, no treatment-related clinical signs or mortality were noted, but slight changes in body weight gain and food consumption were observed in the top-dose group. The parameters of fertility and reproduction showed no significant intergroup differences. At necropsy, slightly higher organ weights (liver, kidney, adrenal gland) and a marginal increase in the incidence and severity of renal tubular basophilia were observed. The F1 litter sizes were not affected by treatment and the average pup weight at birth was similar in all groups. The body weight development of the topdose-group pups was significantly reduced during the lactation period, but the pups reached all milestones of physio­ logical development at the same time as the untreated animals. The parameters of fertility and reproduction were similar in treated and untreated groups. At necropsy, increased liver weights were noted, but the histopathological examination revealed no changes of toxicological significance. In the F2 litters of the top-dose group, the mean pup weight at birth was slightly lower than the control value. All litters reached the milestones of physical development at similar time points. In conclusion, the results of this study gave no indication of effects on reproduction or fertility.

89.5.7  Neurotoxic Effects The neurotoxic potential of Cyprodinil was investigated in the following studies: Rat, single oral dose Rat, 90-day feeding

l l

89.5.7.1  Rat, Single Oral Dose Groups of 10 male and female Tif:RAIf (SPF) rats received a single oral dose of 0, 200, 600, or 2000 mg/kg. The study included a functional observational battery that covered central nervous system (CNS) activity, CNS excitation,

Hayes’ Handbook of Pesticide Toxicology

and sensorimotor, autonomic, and physiological functions. Neurological examinations covered sensorimotor functions, autonomic functions, and sensorimotor coordination. Motor activity was assessed and neuropathological examinations included different areas of the brain, the spinal cord, peripheral nerves, and muscles. The time of peak effect, determined in a range-finding test, was found to be approximately 2 h after treatment. There was no effect of treatment on mortality, body weight, or food consumption. Observations and functional tests showed relevant changes mainly at the time of peak effect. In females of the intermediate and high-dose groups, reduced activity, hunched posture, piloerection, and decreased responsiveness to sensory stimuli were observed; hunched posture was also seen in a few low-dose females. In females of the two higher-dose groups, signs lasted up to test day 4 and were considered to indicate toxicity. In addition, a dose-related decrease in body temperature was observed. Changes in motor activity parameters were limited to the time of peak effect. In high-dose males and in females of the two higher-dose groups, horizontal and vertical activity parameters were reduced. Macroscopic and microscopic examinations of the multiple areas of the central and peripheral nervous system, the eyes, optic nerves, and skeletal muscle of the male and female control and high-dose animals did not reveal treatment-related neuropathologic changes. The results of this study gave no indication of neurotoxic effects.

89.5.7.2  Rat, 90-Day Feeding Groups of 10 male and female Tif:RAIf (SPF) rats received Cyprodinil at dietary concentrations of 0, 80, 800, and 8000 ppm for 90 days. The study included a functional observational battery, covering CNS activity, CNS excitation, and sensorimotor, autonomic, and physiological functions. Neurological examinations covered sensorimotor functions, autonomic functions, and sensorimotor coord­ ination. Motor activity was assessed and neuropathological examinations included different areas of the brain, the spinal cord, peripheral nerves, and muscles. There was no effect on mortality and clinical signs. In high-dose animals, a moderately reduced body weight gain was seen throughout the treatment period. In high-dose animals, absolute and relative liver weights were increased in both sexes and kidney weights were elevated in females. Microscopic examination revealed hepatocellular hyper­trophy in the liver of high-dose animals. In addition, chronic tubular lesions combined with tubular casts and single cystic changes in the kidneys, and hypertrophy of the follicular epithelial cells in the thyroid gland were seen in these animals. Observations and functional tests showed no effect of toxicological significance and no treatment-related effects on the different motor activity parameters. Neuropathologic examination of the eyes, optic nerves, and multiple areas

Chapter | 89  Cyprodinil: A Fungicide of the Anilinopyrimidine Class

of the central and peripheral nervous system of the male and female control and high-dose animals revealed no treatment-related neuropathologic changes. The results of this study gave no indication of neurotoxic effects.

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89.6  Mutagenicity

89.5.8.4  In Vitro, Autonomic Nervous System

The mutagenic potential of Cyprodinil was investigated in five independent studies that covered different endpoints in eukaryotes and prokaryotes in vivo and in vitro. No induction of back mutations was noted in four strains of Salmonella typhimurium (TA 98, TA 100, TA 1535, and TA 1537) or in Escherichia coli WP2 uvrA. The compound was tested in the presence and in the absence of an extrinsic metabolic activation system (rat liver S9 fraction), covering a concentration range of 20–5000 g/plate. The induction of gene mutations was further investigated at the hprt gene of Chinese hamster V79 cells. Concentrations up to the limit of toxicity of 96 and 30 g/ ml were tested in the presence and in the absence of rat liver S9 fraction, respectively. No increased incidence of gene mutations was detected in this in vitro system. Chinese hamster CHO cells were used for the detection of chromosome aberrations in vitro. Three experiments in the presence and in the absence of rat liver S9 fraction were performed by applying different exposure scenarios. Concentrations higher than 50 g/ml could not be investigated due to cytotoxicity. In none of the experiments was an increased incidence of metaphases containing chromosome aberrations seen. Primary cultures of rat hepatocytes were treated with Cyprodinil at concentrations up to 80 g/ml to investigate DNA damaging effects. Higher concentrations caused cytotoxicity. None of the tested concentrations caused enhanced unscheduled DNA synthesis, which would be indicative of the DNA damaging activity of the compound. The formation of micronuclei was investigated in bone marrow cells of Tif:MAGf (SPF) mice. Single oral doses of up to 5000 mg/kg body weight were administered 16, 24, and 48 h before preparation of bone marrow. No cytotoxicity was observed at any dose level and the incidence of micronucleated polychromatic erythrocytes was not affected. Therefore, there is no evidence for a clastogenic or aneugenic activity of Cyprodinil in somatic cells in vivo.

Cyprodinil inhibited the induced contractions of isolated guinea pig ileum at concentrations of 10 μg/ml. No effects were noted at lower concentrations.

89.7  Toxicity to humans

89.5.8  Pharmacological Effects General pharmacological studies were conducted with male ICR mice, male Wistar rats, and male Hartley guinea pigs. Single-dose levels of 0, 150, 500, 1500, and 5000 mg/ kg Cyprodinil were orally administered. In the in vitro experiments, the test material was dissolved in ethanol and applied at concentrations of 0, 0.1, 1.0, and 10 g/ml.

89.5.8.1  Mouse, General Behavior   and CNS Tests No animals died at any dose level. At 1500 mg/kg, slightly decreased spontaneous motor activity, slightly dilated pupil size, and slightly narrowed palpebral opening were observed within the first few hours. All signs disappeared by 6 h after treatment. At 5000 mg/kg, slight piloerection and abnormalities of body and limb position were seen in addition. All signs disappeared within 24 h in this dose group. Doses of 1500 mg/kg or higher prolonged hexobarbital-induced sleeping time. No effects on tonic extensor, clonic convulsions, or coma in electrically stimulated mice were noted at any dose level.

89.5.8.2  Rat, Body Temperature Body temperature was not affected in any dosage group.

89.5.8.3  Rat, Cardiovascular System Treatment with Cyprodinil had no effect on the systolic blood pressure. At 5000 mg/kg, the heart rate was significantly decreased at 1 h after dosing.

89.5.8.5  Mouse, Gastrointestinal System No effect was seen on intestinal transport at any dose level.

89.5.8.6  Mouse, Skeletal Muscle System No effect was seen in the traction test at any dose level.

89.5.8.7  Rat, Hematology No effects on prothrombin time or activated partial thromboplastin time were seen at any dose level. No hemolytic effects were noted.

89.7.1  Direct Observations and Health Records Three cases of moderate, reversible local irritation (erythema, swelling of eyelids) occurred among laboratory personnel during formulation development. No further complaints were noted.

89.7.2  Diagnosis of Poisoning In animal studies, symptoms of acute intoxication were unspecific and transient only. The same can be expected for

Hayes’ Handbook of Pesticide Toxicology

1912

humans. However, no case of intoxication with Cyprodinil has yet been observed.

89.7.3  Sensitization Observations No cases of skin sensitization were recorded in humans.

89.7.4  Proposed Treatment Whereas no specific antidote is known, symptomatic therapy is to be given to persons who show symptoms after exposure to Cyprodinil.

Conclusion Cyprodinil is an anilinopyrimidine fungicide highly effective against a broad range of fungi. The toxicological properties of Cyprodinil have been extensively evaluated in several animal species for acute, subchronic, and chronic toxicity, carcinogenicity, reproductive and developmental toxicity, neurotoxicity, and genotoxicity. The effects of Cyprodinil on liver enzymes and liver and plasma lipids have been studied in rats. Toxicokinetics and metabolism were investigated in a number of animal species. Absorption, distribution, metabolism, and excretion of Cyprodinil have been studied in rats, goats, and hens. In rats, Cyprodinil was rapidly absorbed from the gastrointestinal tract. Elimination occurred via bile, urine, and feces and the elimination was essentially independent of sex, dose level, and pretreatment. Residues in tissues were generally low and no accumulation was observed. Cyprodinil was almost completely metabolized and only minor amounts were excreted unchanged in feces. A small quantitative sex difference in the ratios of two urine metabolites was observed. In goats, Cyprodinil was absorbed to a lesser extent and more slowly than in the rat. It was excreted mainly via feces and urine, with only minimal amounts detectable in milk. Tissue residues of Cyprodinil were generally low. In hens, Cyprodinil was rapidly and completely eliminated, and residues in eggs and tissues were low. The metabolic pathway of Cyprodinil in hens and goats was similar to the pathway in rats. Cyprodinil has low acute toxicity characteristics and is not irritating to skin and eye but may cause skin sensitization. Short-term dietary administration of Cyprodinil to rats and dogs resulted in reduced food consumption and reduced body weight gain at the top dose. The primary target organs were the liver, kidney, and thyroid in rats, and the liver in dogs and mice. In rats, increased liver weights, hepatocellular hypertrophy and minimally increased activity of hepatic enzymes at top doses could be interpreted as an adaption response. The slightly increased plasma concentrations of cholesterol and phospholipids seen in the 90-day rat study

at the two top-dose levels confirmed that Cyprodinil, like other anilinopyrimidines, interacts with lipid homeostasis. The exact mechanism by which these compounds interfere with blood cholesterol and phospholipid synthesis is not known. These effects were not observed at lower dose levels. No evidence of a carcinogenic potential was seen after chronic administration of Cyprodinil to rats and mice. Increased liver weights were observed in top-dose-level male rats and slightly increased liver and kidney weights were seen in top-dose-group mice of both sexes. No adverse effects on reproduction or fertility were seen in a rat two-generation study. Treatment of pregnant rats and rabbits with Cyprodinil gave no indication of teratogenic effects. Cyprodinil was found to have no mutagenic potential when tested in a battery of in-vitro and in-vivo tests. Acute and subchronic neurotoxicity studies in rats showed no evidence for a neurotoxic potential of Cyprodinil. Despite close structural relationship, the anilinopyrimidines Cyprodinil and Mepanipyrim show different effects in the subchronic rat and dog studies. While Cyprodinil caused elevated blood concentrations of cholesterol and phospholipids in rats after 90 days of treatment, administration of Mepanipyrim for 3 weeks caused decreased blood cholesterol, triglyceride, and phospholipid levels in male rats. In addition, the hepatocellular fatty vacuolation seen after 13 weeks of Mepanipyrim treatment did not occur with Cyprodinil. The lipofuscin deposition in Kupffer cells and hepatocytes observed after subchronic administration of Mepanipyrim to dogs was also not observed with Cyprodinil.

References Carroll, R. M., and Feldman, E. B. (1989). Lipids and lipoproteins. In “The Clinical Chemistry of Laboratory Animals” (W. F. Loeb and F. W. Quimby, eds.), pp. 95–116. Pergamon, New York. Fritz, R., Lanen, C., Colas, V., and Leroux, P. (1997). Inhibition of methionine biosynthesis in Botrytis cinerea by the anilinopyrimidine fungicide pyrimethanil. Pestic. Sci. 49, 40–46. Heye, U. J., Speich, J., Siegle, H., Steinemann, A., Forster, B., KnaufBeiter, G., Herzog, J., and Hubele, A. (1994). CGA 219417: A novel broad-spectrum fungicide. Crop. Protection 13, 541–549. Hurley, P. M. (1998). Mode of carcinogenic action of pesticides inducing thyroid follicular cell tumors in rodents. Environ. Health Perspect. 106, 437–445. Lake, B. G., and Lewis, D. F. V. (1993). Structure-activity relationships for chemically induced peroxisome proliferation in mammalian liver. In “Peroxisomes: Biology and Importance in Toxicology and Medicine” (G. Gibson and B. Lake, eds.). Taylor and Francis, London. Masner, P., Muster, P., and Schmid, J. (1994). Possible methionine biosynthesis inhibition by pyrimidinamine fungicides. Pestic. Sci. 42, 163–166. Milling, R. J., and Richardson, C. J. (1995). Mode of action of the anilinopyrimidine fungicide pyrimethanil. 2. Effects on enzyme secretion in Botrytis cinerea. Pestic. Sci. 45, 43–48.

Chapter | 89  Cyprodinil: A Fungicide of the Anilinopyrimidine Class

Miura, I., Kamakura, T., Maeno, S., Hayashi, S., and Yamaguchi, I. (1994). Inhibition of enzyme secretion in plant pathogens by mepanipyrim, a novel fungicide. Pestic. Biochem. Physiol. 48, 222–228. Terada, M., Mizuhashi, F., Tomita, T., Inoue, H., and Murata, K. (1998a). Mepanipyrim induced fatty liver in rats but not in mice and dogs. J. Toxicol. Sci. 23, 223–234. Terada, M., Mizuhashi, F., Tomita, T., and Murata, K. (1998b). Effects of mepanipyrim on lipid metabolism in rats. J. Toxicol. Sci. 23, 235–241.

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Terada, M., Mizuhashi, F., Murata, K., and Tomita, T. (1999). Mepanipyrim, a new fungicide, inhibits intracellular transport of very low density lipoprotein in rat hepatocytes. Toxicol. Appl. Pharmacol. 154, 1–11. Whitlock, J. P., and Denison, M. S. (1994). Induction of cytochrome P450 enzymes that metabolize xenobiotics. In “Cytochrome P450: Structure, Mechanism, and Biochemistry” (R. Ortiz de Montellano, ed.) 2nd ed, pp. 367–390. Plenum, New York. Yamagata, S. (1989). Roles of O-acetyl-L-homoserine sulfhydrylase in microorganisms. Biochimie 71, 1125–1143.

Chapter 90

Captan and Folpet Elliot B. Gordon Elliot Gordon Consulting, LLC

90.1  Introduction Captan and folpet are fungicides that have been in use for over 55 years. During this period there have been no reports of systemic toxicity, but there has been a low incidence of skin sensitization. This record of safe use is consistent with these compounds’ chemical and physical properties. Adverse findings in laboratory test systems consist of mutagenicity, carcinogenicity, sensitization, and eye irritation. Until 2004, the U.S. EPA classified both compounds as “probable human carcinogens.” In the 2004, U.S. EPA reclassified captan as “not likely” based on exposure levels from registered uses. The potential for eye irritation is the basis for farm worker re-entry restrictions. The U.S. EPA is currently re-evaluating folpet for human carcinogenicity classification; it is expected that a similar “not likely” classification be established. Both captan and folpet irritate the gastrointestinal tract of mice when administered at high dietary doses. Adenomas and adenocarcinomas develop primarily in the duodenum, following prolonged compensatory proliferation of duodenal crypt cells following damage to villi. These fungicides are mutagenic when tested in vitro (e.g., the Ames point mutation assay) but negative when tested in vivo (e.g., the micronucleus assay). This paradox reflects the rapid degradation of solubilized compounds in blood: captan degrades with a half-life of less than 1 s; folpet degrades with a half-life of less than 5 s. The margins of exposure (MOEs) based on the no-observed-effect levels (NOELs) for gastrointestinal irritation and subsequent tumor formation is at least 1,000,000 based on estimated human exposure. In practical terms, these high MOEs mean neither captan nor folpet pose a risk for tumors in humans. Draize rabbit studies show that these compounds are severe ocular irritants. Extensive experience, however, particularly with agricultural workers engaged in re-entry operations, has shown that this laboratory phenomenon is not Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

predictive of human experience. Eye protection is indicated for operators who mix, load, and apply these fungicides. Captan and folpet remain efficacious fungicides that present low manageable risks to agricultural workers and consumers.

90.1.1  Overview Captan and folpet are broad-spectrum protectant fungicides. Their mode of action centers on their reaction with thiols. These compounds along with a third, captafol, are collectively called chloroalkylthio fungicides due to the presence of side chains that contain chlorine, carbon, and sulfur. Of the chloroalkylthio fungicides, captan and folpet predominate in agronomic practice today; captafol registrations in the United States were withdrawn in 1988. Related compounds associated with this fungicide class, but not registered in the United States, are dichlofluanid and tolylfluanid. These later two compounds have a fluorine atom substituted for one of the terminal chlorine atoms. Early investigations on captan and folpet focused on their mutagenicity. These assays, conducted in vitro, showed both to be mutagenic. Citing this mutagenicity, regulators ascribed a genotoxic basis to the development of mouse duodenal tumors. This, in turn, led to an initial cancer risk assessment based on a linear low-dose extra­polation. Developmental toxicity studies of folpet were initiated following the perceived association of folpet’s phthalimide moiety with the human teratogen S-thalidomide; these structures have since been shown to be toxicologically unrelated. Captan and folpet show developmental toxicity at maternally toxic doses; neither compound is a frank teratogen. A number of reviews have addressed the toxicology of the chloroalkylthio fungicides (Ecobichon, 1996; Edwards et al., 1991; Elder, 1989; IARC, 1983; Saunders and Harper, 1994; Trochimowicz et al., 2001; U.S. EPA, 1975). 1915

1916

The relatively stable degradates of captan (tetrahydroph­ thalimide, THPI) and folpet (phthalimide, PI) have low toxicity; thus, there is low risk of systemic toxicity to farm workers who mix, load, or apply these fungicides.

90.1.2  History and Use Captan was first registered in the United States on March 8, 1949, as a fruit tree spray (NPIRS, 1999a) and its properties were described in 1953 (Kittleson, 1953). This compound proved extremely efficacious, spurring chemists to turn out a series of analogues in an attempt to capitalize on the fungicidal properties of the trichloromethylthio moiety (Horsfall and Rich, 1957; Kittleson, 1953; Lukens, 1966). Folpet was synthesized after captan; captafol was the last to be developed. As preventative fungicides, they are efficacious when applied prior to the establishment of pathogenic fungi. Captan and folpet are often used in integrated pest management (IPM) programs in conjunction with other fungicides. Registrations cover both agricultural and industrial uses (NPIRS, 1999b; U.S. EPA, 1985b). Captan is also efficacious as a bacteriostat in cosmetics (Elder, 1989). The U.S. EPA issued registration standards for captan (U.S. EPA, 1986b), folpet (U.S. EPA, 1987), and captafol (U.S. EPA, 1984a). A special review for captafol (U.S. EPA, 1985a) concluded in 1988 with the voluntary withdrawal of all registrations. A special review for captan was completed in 1989 with the issuance of Position Document 4 (U.S. EPA, 1989). Reregistration Eligibility Decision documents (REDs) have now been promulgated for captan and folpet (U.S. EPA, 1999a,b).

90.1.3  Toxicological Overview The principle chemical reaction that governs the toxicity of captan and folpet is their rapid reaction with thiol groups (i.e., sulfhydryl, -SH groups). This reaction results in degradation of the parent compound. Thiophosgene is a key degradation product that also reacts with thiols as well as other functional groups. Thiophosgene is more reactive than captan or folpet as reflected by its half-life in blood of 0.6 s (Arndt and Dohn, 2004). The net result of these chemical interactions is that both captan and folpet elicit primary toxicological effects locally at the site of initial contact. In mice, dietary exposure results in local irritation of the gastrointestinal tract, predominantly in the duodenum. It is primarily at this site that tumors develop. Continued administration of high doses will lead to secondary effects such as decreased body weight gain or growth retardation of fetuses in developmental studies. Pursuant to the Food Quality Protection Act and subsequent guidance by the EPA, captan and folpet have been found to share a common mechanism of toxicity with regard to the development of duodenal tumors in mice (see

Hayes’ Handbook of Pesticide Toxicology

discussion in Section 90.4). This finding suggests that both compounds be considered for cumulative risk assessment; it also affords the toxicologist the opportunity to combine mechanistic data for each into one unified database. Captafol, in contrast, was found not to share a common mechanism of toxicity with captan or folpet with regard to elicitation of systemic tumors. The FQPA requires consideration of the special sensitivities to infants and children, and the potential for endocrine disruption (Gabriel, 2005; U.S. EPA, 2008). There is no indication that captan or folpet shows disproportionate toxicity to infants or children. The mechanism by which these compounds exert their toxicity would make such a distinction unexpected. The EPA has recognized this for captan but currently has assigned an additional threefold safety factor for folpet (see Section 90.3.4; U.S. EPA, 1999a,b). Captan or folpet shows little evidence of being endocrine disruptors; however, captan has been “strongly suspected” of acting as an antiestrogen (Okubo et al., 2004). A Tier I endocrine screening program is currently under development (Gabriel, 2005; U.S. EPA, 2008). Captan and folpet induce mutagenic and clastogenic effects in a variety of in vitro assays. Mutagenic effects in vivo, however, do not occur. This paradox is explained by the extremely rapid degradation of these compounds in the intact animal. Whereas the delivered dose is negligible, captan and folpet are classic examples of the adage, “the (delivered) dose makes the poison.” Despite the obvious potential for mutagenic events, the dose at sensitive targets in the intact animal, such as cellular DNA, is essentially zero. Although these fungicides have low acute toxicity, their interaction with biological tissues can cause irritation. Persons handling these materials should do so with appropriate respiratory and eye protection. The compounds are not considered reproductive toxins or selective developmental toxins. The appearance of duodenal tumors in mice fed diets admixed with captan or folpet was, until recently, a key toxicological finding central to the regulation of these compounds. Data show that a mode of action based on increased rates of cell proliferation, a threshold phenomenon, accounts for these tumors. This proliferative pressure promotes nascent tumor cells that are normally resident within the duodenal crypt compartment. EPA accepted this mode of action when it reclassified captan (U.S. EPA, 2004). EPA considered comments from the public regarding this reclassification and affirmed the science supporting their decision (Jennings, 2007; Kent, 2006; Koch, 2007). Cancer reclassification of pesticides is now incorporated into the Pesticide Registration Improvement Renewal Act (U.S. EPA, 2007); the work leading to captan’s reclassification helped make these reevaluations routine (Gordon, 2007). Farm workers who handle captan or folpet are not at risk for developing duodenal cancer, as there is no systemic exposure. Persons exposed to residues of captan

Chapter | 90  Captan and Folpet

1917

and/or folpet in their food are not at risk since the margins of safety for both compounds are approximately 1 million (see Section 90.5).

90.2  Physical properties and chemical reactions

but is subordinate to the side chain with regard to the fungicide’s toxicological properties. The phthalimide ring is aromatic and, as such, is a resonance structure; THPI has one double bond between carbons 3 and 4. This hexene imide, unlike phthalimide, is nonplanar (Fickentscher et al., 1977). Chemical

Structure

90.2.1  Overview

O

Cl N—S–C–Cl – –

The toxicology of the chloroalkylthio fungicides is dependent on their physical properties and chemical reactions. The structures of captan and folpet along with typical ring degradates are shown in Figures 90.1 and 90.2. The chemical identity and physical properties are noted in Table 90.1, and the rates of selected chemical reactions are shown in Table 90.2. The characteristic chemical moiety for captan and folpet is the trichloromethylthio side chain that is connected to an imide ring structure by way of a nitrogen-sulfur bond. Captan’s ring is tetrahydrophthalimide (THPI) and folpet’s is phthalimide. This ring imparts certain physical properties to the molecule,

Folpet

Cl

O O Phthalimide (PI)

NH O O NH2

Phthalamic acid

OH O O

Chemical

Structure O

Cl

– – Cl

O

O NH

4,5-cyclohexene-1,2-dicarboximide (THPI)

4,5-dihydroxy-1,2-dicarboximide (4,5-diOH THPI)

Parameter

Captan

Folpet

CAS number

133-06-2

133-07-3

Molecular weight

300.61

296.56

Formula A

C9H8Cl3NO2S

C9H4Cl3NO2S

Formula B

C6H8 (C O)2N— SCCl3

C6H4(C  O)2N— SCCl3

IUPAC name

1,2,3,6-TetrahydroN-(trichloromethyl thio) phthalimide

N-(trichloromethyl thio) phthalmide

CA name

3a,4,7,7aTetrahydro-2[(trichloromethyl) thio]-IH-isoindole1,3(2H)-dione

3-[(Trichloromethyl) thio]-IH-isoindol1,3(2H)-dione

O CNH2

Physical form

Crystals

Crystals

Melting point

178°C

177°C

COOH

Solubility, water

3.3 mg/l at 25°C

1 mg/l at 20°C

Solubility, acetone

3.0 g/100 ml

3.4 g/100 ml

log Kow

2.35

2.85

O NH

O O

O

OH

NH OH

O

OH O NH

7-hydroxy-4,5-cyclohexene-1,2-dicarboximide (ci/trans-3-OH THPI)

O

O 6-hydroxy-4,5-cyclohexene-1,2-dicarboximide (cis/trans-5-OH THPI)

OH NH O

1-amido-2-carboxy-4,5-cyclohexene (cis/trans-THPAM)

6-hydroxy-1-amido-2-carboxy-4,5-cyclohexene (3-OH THP-amic acid)

Figure 90.2  Folpet and its ring metabolites.

Table 90.1  Physical Properties of Captan and Folpet

O 4,5-epoxy-1,2-dicarboximide (THPI expoxide)

OH O CNH2 COOH

Figure 90.1  Captan and its ring metabolites.

OH O

N—S–C–Cl

Captan

OH

Phthalic acid

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Table 90.2  Rates of Chemical Reactions of Captan and Folpet Parameter

Captan

Folpet

Reference

pH 5

18.8 h

2.6 h

Captan: Pack (1987)

pH 7

4.9 h

1.1 h

Folpet: Ruzo and Ewing (1988)

Aqueous hydrolysis

pH 9

8.3 min

1.5 min

Reaction with blood thiols

18 s, 22°C

51 s, 22°C

Captan: Crossley (1967a) Folpet: Crossley (1967b)

Degradation half-life (acid conditions to minimize hydrolysis)

0.97 s, 37°C

4.9 s, 37°C

Gordon et al. (2001)

Decolorization of dithionitrobenzoic acid (DTNB)-thiol complex in blood

1 min

3 min

Liu and Fishbein (1967)

90.2.2  Physical Properties Captan and folpet have similar physical properties. They have low water solubility, low volatility, and melt at approximately the same temperature. Octanol–water coefficients are high for both, although folpet’s Kow is somewhat higher than that of captan.

90.2.3  Chemical Reactions Captan and folpet are unstable in aqueous solution, but the rate of hydrolysis is slow compared with their reaction with thiols. The key to their fungicidal efficacy is the balance between the reactivity of the trichloromethylthio moiety and the stability of the nitrogen–sulfur bond linking this moiety to the imide ring. Analogues with very stable bonds prove to be ineffective fungicides, whereas analogues with bonds that are overly labile degrade spontaneously (Horsfall and Rich, 1957; Lukens, 1966, 1967). The hydrolytic and thiol reactions serve to degrade the parent molecule and thus influence the toxicology outcome by effectively reducing or eliminating exposure.

90.2.3.1  Hydrolysis The rates of aqueous hydrolysis increase in alkaline conditions and are more rapid for folpet than captan at comparable

pH values (Table 90.2). At pH 5, for instance, captan is approximately eight times more stable than folpet; thus, in the acid conditions of the stomach, it would be expected that relatively more folpet degradation products would be present compared with captan. The higher hydrolytic rates for folpet are related to the higher standard free energy of the phthalimide ring structure compared with the THPI ring (Lukens, 1966).

90.2.3.2  Reaction with Thiols The fungicide-thiol reaction has been studied with glutathione (GSH), proteins, and other thiol-containing compounds. In general, the thiol group is oxidized (e.g., GSH → GSSG; cysteine → cystine). Common to both captan and folpet is the generation of thiophosgene during degradation. This chemical entity appears to be a contributing toxicophore in that it rapidly reacts with a variety of functional groups in addition to thiols (Lukens, 1969; Lukens and Sisler, 1958b; Sharma, 1986). A general scheme of degradation for captan and folpet is shown in Figure 90.3. The rate of hydrolysis is faster for folpet than for captan, whereas the reverse is true for thiol-mediated degradation. The reaction of captan and folpet with cysteine results in the formation of thiazolidine-2-thione-4-carboxylic acid (TTCA; Lukens and Sisler, 1958a). This compound is seen in mammalian metabolism studies (DeBaun et al., 1974) and has been suggested for use as a biological marker for human exposure assessment (Krieger and Dinoff, 2000; Krieger and Thongsinthusak, 1993; van Welie et al., 1991). TTCA can now be detected at a level of 40 pmol/ml urine (Amarnath et al., 2001). The fate of captan and folpet in human and rabbit blood has been investigated (Crossley, 1967a,b). Crossley added captan and folpet to human blood and measured the decline of the parent with time and, concurrently, the increase of the imide ring (THPI or phthalimide). At initial concentrations of 1 g/ml, captan degraded with a half-life of 18 s. The degradation of folpet was three times slower, with a half-life of 54 s. By measuring the generation of the imide rings, it was shown that the parent compounds actually degraded rather than complexed with blood constituents. These investigations were carried out at 22°C with unlabeled materials. Subsequent investigations of degradation rates employed radiolabeled captan and folpet and physiological temperatures (37°C). As predicted by the Q10 (Purves et al., 1992), increased temperature results in a higher rate of degradation. The new data demonstrate that at physiological temperatures, captan degrades rapidly in human blood, having a t1/2 of 0.97 s, whereas folpet degrades somewhat slower but still quite rapidly (i.e., t1/2  4.9 s; Gordon et al., 2001). These data, which demonstrate the rapid degradation of the two compounds, are of course a critical component in any exposure assessment and risk

Chapter | 90  Captan and Folpet

Reaction with thiols

1919

R—SCCl3 Captan/Folpet

Hydrolysis

2R’SH

R’SSR’

H2O

HCl

THPI (captan) PI (folpet) HCl

S C Cl Cl Thiosphosgene Reaction with thiols

Hydrolysis

which such effects are induced has not been elucidated. When 35S-captan was incubated with calf thymus DNA in buffer at pH 7.5 or 9.0, binding of appreciable amounts of radioactivity could not be demonstrated (Couch and Siegel, 1977). Captan reacts with guanine in vitro to produce 7-(trichloromethylsulfenyl) guanine (Elder, 1989), as reported by FAO/WHO (1990). In vivo DNA binding studies are discussed in Section 90.3.5.5.

90.2.3.5  Miscellaneous Chemical Reactions

evaluation modeling. With oral exposure, it is unlikely that captan, folpet, or thiophosgene (with a half-life of 0.6 s in blood) would survive long enough to reach systemic targets such as the liver, uterus, or testes. With dermal exposure and subsequent low absorption, captan will be eliminated in less than 7 s and folpet in less than 35 s. This determination reflects the kinetics noting compounds are essentially gone in seven half-lives (Medinsky and Klaassen, 1996).

Because captan and folpet are reactive, there are unlimited opportunities for chemical reactions in isolation. Absent data that indicate exposure in vivo or relevance to the intact animal, these observations remain ancillary, reflecting their chemical reactivity, but having little bearing on mammalian toxicity and human risk assessment. Captan and folpet react with p-nitrothiophenol via the thiol group (Liu and Fishbein, 1967). This differential rate of reaction was measured at 25°C and was 1.9  104 l/(mol min) for captan and 1.5  104 l/(mol min) for folpet. Other effects include the inhibition of Escherichia coli RNA polymerase (Elder, 1989), the inhibition of RNA synthesis by intact bovine nuclei (Elder, 1989), the inhibition of microsomal cytochrome P450 benzphetamine N-demethylase and aniline hydroxylase after intraperitoneal dosing (Dalvi, 1988, 1989), the inhibition of the Ca2 transport ATPase in human erythrocytes (Janik, 1986), and the inhibition of oxidative phosphorylation in rat liver mitochondria, correlated to mitochondrial swelling (Elder, 1989). Captan also disrupted the differentiation of cultured cells from the midbrains and limb buds of 34–36 somite rat embryos in vitro (Flint and Ortaon, 1984) and inhibited the attachment of tumor cells to polyethylene disks that were coated with concanavalin (Braun and Horowicz, 1983).

90.2.3.3  Reaction with Proteins

90.2.3.6  Thiophosgene

Investigations on the effects of captan and folpet with proteins generally have been carried out in vitro. Such studies identify potential interactions that may occur in the living animal; however, for captan and folpet, the rapid degradation of reactive species and the resultant limitation in exposure prevent many of these reactions from resulting in in vivo toxicological phenomena. Folpet reacts with thiol-containing proteins (e.g., glyceraldehyde 3-phosphate; Siegel, 1971a), non-thiol-containing proteins (e.g., -chymotrypsin; Siegel, 1971b), and nuclear histones (Couch and Siegel, 1972, 1977). These reactions are often pH-dependent.

Thiophosgene (CAS 463–71–8) is a very short-lived compound that has a broad spectrum of reactions with a variety of functional groups (Sharma, 1978, 1986). Although this compound hydrolyzes at a slower rate than its oxygen analogue, phosgene, the rate is sufficient to eliminate mutagenic activity in Salmonella typhimurium TA 100 when dimethyl sulfoxide (DMSO) is the solvent (Schuphan et al., 1981). Thiophosgene is a toxicophore of captan and folpet, although its role in their fungicidal properties has been questioned (Lien, 1969). It is volatile and reacts with water to form carbonyl sulfide (COS) and two molecules of hydrogen chloride. The carbonyl sulfide then reacts with another water molecule to form hydrogen sulfide and carbon dioxide (Figure 90.3). Thiophosgene has two reactive sites associated with the carbon atom. Whereas both chlorine atoms are electronegative, the carbon atom becomes positively charged, thus creating an electrophile. The reaction with cysteine is shown in Figure 90.4.

TTCA Cysteine + 2HCl

O

2RSH

2H2O 2HCl + CO2 + H2S

SO3=

RSCSR + 2HCl RSR +CS2

Reaction with sulfite DMS-Acid [O] DMS-O

Figure 90.3  General degradation scheme. For captan, R  THPI (Tetrahydrophthalimide); for folpet, R  PI (phthalimide); TTCA: thiazolidine-2-thione-4-carboxylic acid; DMS-Acid: dithio-bis-methanesulphonic acid; DMS-O: monosulfoxide of dithio-bis-methanesulphonic acid.

90.2.3.4  Reaction with DNA The reactions of captan and folpet with DNA are not well characterized. Captan and folpet induce point mutations and clastogenic changes in vitro, but the mechanism by

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– –

O

NH2 Cysteine

+

S Thiophosgene

– –

C

C–OH

S NH C S

+ 2HCl

–

OH

C– C

–

–

HS

– – –

Cl Cl

– – –

O

Thiazolidine-2-thione-4carboxylic Acid (TTCA)

Figure 90.4  Reaction of cysteine with thiophosgene.

When introduced into human blood, the half-life of thiophosgene is 0.6 s, a value that reflects its high reactivity (Arndt and Dohn, 2004),

90.2.4  Metabolism The fate of captan and folpet in mammalian systems is determined by an amalgam of nonenzymatic chemical reactions with thiols and subsequent enzyme-mediated metabolism that predominately involve the generation of ring metabolites. In the intestine, both hydrolysis and thiolmediated reactions occur. The rate of hydrolysis is particularly sensitive to pH, and the transition from the acid environment of the stomach to the neutral or basic conditions of the duodenum promotes the hydrolytic breakdown of these materials. These fungicides undergo a similar pattern of degradation (Figure 90.3). The side chain is either fully mineralized or forms by-products such as TTCA with cysteine. The respective imides, THPI and phthalimide, are initially formed either through hydrolysis or through reaction with thiols. These are subsequently metabolized to secondary products; the THPI hexene structure of captan is more extensively metabolized than the phthalimide structure of folpet (Figures 90.1 and 90.2).

90.2.4.1  Rat Metabolism Captan and folpet are rapidly eliminated when administered either orally or intraperitoneally. Multiple doses of captan or folpet do not alter subsequent excretory patterns, suggesting that liver enzymes are not induced by repeated exposure. This finding was expected because the parent molecules are not likely to reach the liver. There is no sex difference in the way these fungicides are metabolized. A 10-mg/kg dose of ring-labeled captan is rapidly excreted in the urine. After 24 h, approximately 75% of the administered dose is excreted in the urine and 6.5% is excreted in the feces. Nearly all radioactivity is excreted by 36 h (Trivedi, 1990a). Fourteen repeated single doses of 10 mg/kg followed by a dose of radiolabeled captan produced a similar excretory profile (Bratt, 1990). A dose of 6 mg/kg 35Scaptan given intraperitoneally to male rats was effectively eliminated within 72 h (Couch et al., 1977).

Captan at a 500-mg/kg dose resulted in a similar profile except that relatively more material was excreted via the feces. In 96 h, 68.8 and 23.1% was excreted via the urine and feces in males and 73.4 and 25.0% was excreted, respectively, in females (Trivedi, 1990b). Folpet demonstrates a similar pattern. Administration of 10 mg/kg results in approximately 96% of the radioactivity being excreted by 24 h (90% in urine; 6% in feces). With doses 50 times higher, only approximately 69% of the administered dose is cleared by 24 h (47% in urine; 22% in feces). The imide ring is relatively stable and is excreted along with additional ring metabolites. The side chain is unstable and reacts with thiols to form mineralized products such as CO2, HCl, and H2S. In addition, products of the reaction also include TTCA, dithiobis (methanesulfonic acid) and its disulfide monoxide derivative. The reactions of captan and folpet are identical with regard to the -[trichloromethylthio] side-chain reactions. The THPI generated from captan is more easily metabolized than the phthalimide from folpet. This is due to the carbonyl groups of captan that draw electrons away from the hexene double bond, creating a  charge at this site, thereby promoting substitutions. Administration of phthalimide results in metabolism to phthalamic acid (79%, in females) and phthalic acid (7%). Less than 1% of the original phthalimide is recovered in the urine (Chasseaud et al., 1974). Phthalamic acid accounts for 80% of the original dose when 14C-[carbonyl] folpet is given to rats (Chasseaud, 1980).

90.2.4.2  Effect on Glutathione Levels in the Duodenum Sulfhydryl groups are intimately involved with the degradation of captan and folpet. It is therefore of interest to see their effect on GSH. Swiss Webster mice fed captan at 4000, 8000, and 16,000 ppm for 35 days had GSH levels (“soluble thiols”) elevated by day 1 (Miaullis et al., 1980). The percent increase over controls ranged from 146 to 227%. Gavage treatment at a relatively high dose of captan (2000 mg/kg) induced an increase in GSH levels that was observable within 2 h of treatment, whereas a smaller dose (20 mg/kg) induced a measurable increase at 4 h (Katz et al., 1982; Sauerhoff et al., 1982). Folpet-induced increased GSH levels were demonstrated after both dietary administration and gavage (Chasseaud et al., 1991). Folpet, administered by gavage (7.6, 72, and 668 mg/kg), initially induced a decrease (30 and 60 min), which subsequently rebounded to a higher than normal level. The decrease was statistically significant for the 72-mg/kg dose: levels of 76, 54, 82, 155, and 130% (of control values) were observed at 0.5, 1, 2, 6, and 24 h, respectively. This rebound effect was also seen at 668 mg/kg: 72, 52, 94, 143, and 178% at the same time periods. Diethylmaleate produced a similar pattern of GSH loss and rebound. These data demonstrate that captan and folpet

Chapter | 90  Captan and Folpet

cause an initial lowering of GSH levels followed by an increase due to a homeostatic rebound. The generative process for GSH exceeds the loss, and a steady state of higher GSH levels is quickly reached.

90.2.4.3  Goat Metabolism Ring-labeled (14C)captan was administered to goats in gelatin capsules three times per day for 4 days. The total daily dose equaled approximately 50 ppm (Cheng, 1980). Most of the radioactivity was excreted in the urine, and the next highest excretion was via the feces. Five biochemical reactions were noted from this study: 1. Cleavage of the N—S linkage in the captan molecule to form THPI, either by hydrolysis or reaction with SH compounds 2. Ring hydroxylation of THPI to form 3-OH THPI 3. Isomerization of the 3-OH THPI to form 5-OH THPI 4. Epoxidation of THPI to form THPI-epoxide, which is subsequently hydrolyzed to form 4,5-diOH HHPI 5. Hydrolysis of THPI and/or its hydroxylated derivatives to form their corresponding THP-amic acid derivatives When radiolabeled folpet [14C-labeled on the trichloromethylthio side chain (TCM)] was administered via capsules to goats, radioactivity was recovered as expired 14 CO2 (31%; reflecting the breakdown of the TCM), in the feces (21%), and in the urine (10%; Corden, 1997). Recovered 14CO2 from the expired air (ca. 31%) reflected the breakdown of the trichloromethylthio moiety to CO2. Similarly, administration of ring-labeled folpet showed the same excretion pattern as discussed for captan [i.e., more excretion via the urine (58%) than the feces (35%), with the major urinary metabolite being phthalamic acid (49% of the dose)].

90.2.4.4  Hen Metabolism Laying hens metabolize captan in a similar way as mammals (Daun, 1988a,b). When tagged with 14C on the imide ring, the identified metabolites were THPI (15.8–68.9% of the tissue radioactivity), and 3-OH, and 5-OH-THPI, which represented 2.4–26% of the radioactivity. When tagged with 14C on the trichloromethylthio moiety TTCA, dithiobis-methanesulfonic acid and its monosulfoxide were seen. Parent captan was not present in the eggs or tissues.

90.2.4.5  Dermal Absorption Dermal absorption of captan and folpet, as subsequently noted, is estimated at no greater than 0.5% per hour. Captan penetration of skin has been measured in vitro, comparing rat with human, and in vivo, using rats. The rat study measured absorption at 1, 2, 4, and 8 h at 19.4 g/cm2 (0.5 mg/kg) and 194 g/cm2 (5 mg/kg; Adir et al., 1982),

1921

and the data have been interpreted as indicating from 0.4% per hour absorption (Ghali, 1997) to 1.5% per hour (11.7% per 8 h; Thongsinthusak et al., 1999). The 11.7% per 8-h rate has been noted as overly conservative because the test sites were not occluded, allowing contamination of the urine and feces samples. Additionally, the absorption rate did not consider the difference between rat and human skin permeability (Fletcher et al., 1995). The in vitro rat/human comparisons showed that human skin was consistently less permeable than rat skin, but that the ratio of permeability was partly dependent on the concentration of captan applied and the solvent used. A study with 14C-folpet 50 WP in rats indicated a systemic absorption of 0.27% per hour (6.5% in 24 h; Wilson and Wright, 1990). This calculation was based on a least square analysis of 24-h urinary excretion at dose levels of 49, 460, and 4800 g/cm2 (13.2, 3.5, and 1.3%, respectively), matching the excretion to the approximate dermal exposure of 2400 g/cm2, based on the 50-WP formulation concentration. There was rapid uptake of folpet into the skin and 95 and 94% for dose levels 4800 and 460 g/cm2, respectively, were retained there at 24 h. The amount still in the skin after 24 h for the low dose was approximately 85% of the applied dose. A comparison with ring-labeled captan and sidechain-labeled folpet showed no difference in absorption between adult and young Fischer 344 rats, but a lesser amount of folpet was absorbed compared with captan (Shah et al., 1987). At 0.54 and 2.68 m/cm2, captan penetration in adult rats was 3.7 and 3.6%, whereas folpet penetration was 2.7 and 1.1%, respectively. These data were interpreted by the U.S. EPA to suggest 0.4% per hour (4.29% in 10 h) dermal absorption for folpet (Ghali, 1997; Levy et al., 1997). These data show high dermal adsorption, but low penetration rates for captan and folpet. There are differing interpretations of these data, but it is reasonable to conclude that the hourly dermal penetration rate is no greater than 0.5%. The normal sloughing of the stratum corneum serves to deplete the amount available for absorption.

90.2.4.6  Human Metabolism Humans appear to metabolize captan in a similar manner to other mammals (Krieger and Thongsinthusak, 1993). Both THPI and TTCA have been recovered after oral and dermal dosing. Comparable human studies with folpet have not been conducted, but are expected to yield similar results but with the urinary excretion of phthalamic acid (major) and phthalimide (minor).

90.2.5  Summary Captan and folpet are extensively altered in mammalian and avian systems through a combination of enzymatic and nonenzymatic chemical reactions. Two complementary

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processes, hydrolysis and thiol interactions, initially split the fungicides into their respective imide rings and trichloromethylthio complexes. Subsequent reactions, some of which may be enzymatic, produce a series of imide-based degradates and thiophosgene-mediated products. The reactions are rapid and nearly all material is eliminated from the animal within 24–48 h; there is no accumulation of either imide or side chain. Urinary metabolites from the rings differ between captan and folpet, but those associated with the trichloromethylthio side chain are common [e.g., TTCA and dithiobis(methanesulfonic acid)]. Captan or folpet do not survive in the systemic circulation, thus limiting their primary effects to areas of initial contact. Due to this rapid elimination, meat, milk, or eggs from livestock that might have consumed feed with residues of captan or folpet present would be devoid of the parent materials.

90.3  Toxicology 90.3.1  Acute Toxicology 90.3.1.1  Overview Both captan and folpet have low acute toxicity, except for the intraperitoneal route (Table 90.3). The reactivity to mucus membranes is high and the severe eye irritant finding is consistent with this property. When single 0.5-g doses are applied to the skin, the results show mild to low irritancy. Results of guinea pig sensitization studies give both positive and negative results; however, experience with handlers of these materials suggest that some persons (estimated at less than 10% of the population) are susceptible to sensitization.

90.3.1.2  Acute Oral Toxicity The low acute oral toxicity of captan and folpet reflects the rapid degradation of the fungicides once ingested and the absence of intact fungicide at sensitive biochemical targets. The LD50 values are above 5 g/kg for both technical and formulated products, placing them in Toxicity Category IV for U.S. EPA regulatory purposes. The captan 50W formulation LD50 is 8.4 g/kg, whereas the comparable folpet formulation LD50 is greater than 10 g/kg (Ben-Dyke et al., 1970). Results from other investigators consistently show LD50 values above 5 g/kg for both compounds (Boyd and Krijnen, 1968; Nelson, 1949). The imide ring degradates of captan and folpet, THPI and phthalimide, respectively, are stable compared with the parent compounds. Accordingly, some measure of their acute toxicity is in order. Both these compounds have low acute oral toxicity in mammals. The LD50 of THPI is 2 g/kg (Cavalli, 1970); for phthalimide it is greater than 8 g/kg (U.S. EPA, 1974). In contrast to mammals, aquatic organisms are particularly sensitive to captan and folpet, and offer a useful test system to compare the toxicity of parent and degradate. Comparative toxicity in trout show THPI to be approximately 3500-fold less toxic than captan (captan LC50  34 ppb versus THPI LC50  120,000 ppb; U.S. EPA, 1999a). A similar comparison for folpet shows a 3267-fold difference (folpet LC50  15 ppb versus phthalimide LC50  49,000 ppb; U.S. EPA, 1999b).

90.3.1.3  Acute Intraperitoneal Toxicity Intraperitoneal administered acute toxicity studies are not generally required for regulatory purposes because this route of entry affords little information for human risk assessment. The intraperitoneal route bypasses the intestine,

Table 90.3  Acute Toxicity and Captan and Folpeta Parameter

Captan

Reference

Folpet

Reference

Oral LD50, rat

5 g/kg

Gaines and Linder (1986)

5 g/kg

Gaines and Linder (1986)

8 g/kg 50W formulation

Ben-Dyke et al. (1970)

10 g/kg 50W formulation

Ben-Dyke et al. (1970)

Intraperitoneal LD50, rat

40 mg/kg, M 35 mg/kg, F

Copley (1985)

48–52.5 mg/kg

Dickhaus and Heisler (1983)

Dermal LD50, rabbit

2000 mg/kg

Thoa and Redden (1995)

5000 mg/kg

Korenaga (1982)

Irritation, eye

Severe

Thoa and Redden (1995)

Severe

0.5, EPA (1987)

Skin

Minimal

U.S. EPA (1975)

Minimal

U.S. EPA (1987)

Inhalation LC50, rat, 4 h exposure

0.72–0.87 mg/l

Thoa and Redden (1995)

1.89 mg/liter

Cracknell (1993)

Sensitization, guinea pig a

M, males; F, females.

Moderate

Thoa and Redden (1995)

Moderate

U.S. EPA (1987)

Chapter | 90  Captan and Folpet

although materials are still primarily absorbed by the portal system (Rozman and Klaassen, 1996). In the case of oral administration of captan or folpet, only trace amounts of parent material would enter the portal system and would be rapidly degraded. Intraperitoneal administration bypasses this degradation and thus affords an observation into the inherent toxicity of the materials. Male and female SPF Wistar rats treated with 92.7% captan administered by injection in 1% methylcellulose had LD50 values approximately 40 mg/kg for males and 35 mg/kg for females (Copley, 1985). This LD50, lower by over 100-fold compared with oral administration, indicates the inherent toxicity of captan. It also demonstrates the effective barrier provided by the intestine. Male and female Wistar rats injected with 87.5% folpet in 1% methylcellulose had a 24-h LD50 of 48.0–52.5 mg/kg. Deaths occurred between 24 h and 7 days, resulting in a 7-day LD50 of 36–40 mg/kg. There was a steep dose– response curve: at 30 mg/kg, one in 10 deaths were seen, but at 60 mg/kg, 10 in 10 deaths occurred by day 7 (Dickhaus and Heisler, 1983). The intraperitoneal LD50 values for captan and folpet are similar and reflect the rule that governs their toxicological profile: hazard in the absence of exposure limits adverse effects.

90.3.1.4  Acute Dermal Toxicity Captan and folpet pose little hazard of acute toxicity from dermal exposure. Limit doses of 2 g/kg are without effect in rabbits (Foster and Morgan, 1984; Gaines and Linder, 1986). A study with Phaltan 50W (50% folpet) showed that the LD50 was greater than 22.6 g/kg (Kay and Calandra, 1960).

90.3.1.5  Eye Irritation Captan and folpet irritate mucus membranes and there is the potential for damage when they contact the eyes. Bioassays, however, vary in their estimation of this hazard. Captan eye irritation studies show variable results. Minimal damage, as noted by no corneal or iris involvement, and low redness and swelling to the eyelids, has been reported (Harris, 1976). Conversely, severe damage, including corneal opacity, has also been reported (Rosenfeld, 1984). Washing the treated eyes after instillation of test material reduces irritation, as was observed in a study that employed a captan 50W formulation (Sauer and Seaman, 1980). An additional study employed a combination formulation that included captan (8%), folpet (44%), and captafol (8%), and showed conjunctival irritation but no corneal involvement (Cisson et al., 1983). Folpet Technical, in unwashed eyes, induced transient corneal opacity that progressed to vascularization of the cornea in two of six rabbits (Dreher, 1992a). A 100-mg instillation of Folpet Technical caused corneal opacity in some unwashed eyes and no opacity in eyes that were

1923

washed 30 s after instillation (Cisson et al., 1982). All eyes returned to normal by 7 days (washed) or 10 days (unwashed). Phaltan 500 Flowable formulation (50% folpet) instilled in rabbit eyes, followed by a 30-s wash after 30 s, resulted in no corneal opacity and minimal redness and swelling (Mercier, 1988). By 72 h, all swelling had subsided, but there was some residual redness and congestion.

90.3.1.6  Skin Irritation Both captan and folpet elicit very little irritation when applied as a single dose to either intact or abraded skin. Captan Technical applied to rabbit skin at 0.5 g showed no redness or edema at either 24 or 48 h for both intact and abraded test sites (Harris, 1976). Folpet Technical applied to rabbit skin at 0.5 g showed no redness or edema at observation periods up to 72 h (Rees, 1993). Doses of folpet as high as 22.6 g/kg produced only transient redness in rabbits (Kay and Calandra, 1960).

90.3.1.7  Acute Inhalation Toxicity Acute toxicity via the inhalation route of exposure varies somewhat with the specific formulation tested. For captan, the 4-h LC50 was 1.21 mg/l for males and 1.05 mg/l for females (Cummins, 1995). An earlier study reported these values as 0.90 mg/l for males and 0.67 mg/l for females (Blagden, 1991). For folpet, the 4-h LC50 for males and females was 1.89 mg/l (95% confidence limits 1.47–2.31 mg/l; Cracknell, 1993). The particle size mass median equivalent aerodynamic diameter was 4.6–5.2 m. Males were slightly more susceptible than females. The mortality for males was 0 in 5, 3 in 5, and 4 in 5 for dose levels 0.80, 1.60, and 1.99 mg/l, respectively. Females at these respective doses showed mortality of 0 in 5, 1 in 5, and 1 in 5.

90.3.1.8  Skin Sensitization Guinea pig bioassays show that both captan and folpet have the potential to induce delayed contact hypersensitivity reactions. Captan Technical was positive in the Magnusson and Kligman maximization guinea pig assays (Dreher, 1992b). Folpet Technical was tested using the Magnusson and Kligman protocol in which a 10% w/v preparation in propylene glycol was injected by the intradermal route along with a 1:1 preparation of Freund’s Complete Adjuvant. Subsequent topical inductions were made with 50% w/v folpet. Challenge with either 10 or 50% w/v folpet in propylene glycol resulted in positive reactions (LSR, 1993).

90.3.1.9  Human Experience Reports of adverse effects are limited to incidences of skin irritation or sensitization reactions (Guo et al., 1996; Peluso et al., 1991; U.S. EPA, 1989). In human patch tests, nine in 205 (4.4%) subjects showed a positive Draize reaction

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(Marzulli and Maibach, 1973), eight in 150 (5.3%) were sensitized by 1% topically applied captan (Jordan and King, 1977), and 24 in 200 were positive to chemicals in the fungicide category, predominantly the thiophthalimides (Lisi et al., 1986). In a patch test for photoallergens four of 12 patients who had positive reactions to a series of allergens reacted to folpet and captan (Mark et al., 1999). Additionally, a case of urticaria was associated with use of a captan-formulated product (Croy, 1973). However, a powder blush that contained 0.3% captan failed to sensitize any of 25 adult volunteers on which it was tested. This was in spite of the study design, which included repeated doses (five consecutive 300-mg induction exposures to the forearms) and occlusion of the application site for 48 h after each dose. The individuals were challenged 10 days later and all individuals were negative (Ivy Research Laboratories, 1981). Captan was noted as “the most common sensitizer” affecting five of 30 fruit and vegetable workers in Hamachi Prades, India (Verma et al., 2007). Regular exposure to captan, formulated in a shampoo at 7%, appears to be well tolerated (Guo, 2001). Although there is potential for eye irritation based on laboratory studies, there are no reports in the literature of adverse effects (NLM, 2001). Likewise, eye injuries in agricultural workers who carry out re-entry activities do not appear to be problematic (Krieger, personal communication). In an apparent suicide attempt, a 17-year-old ingested 7.5 g captan 50 WP. There were increases in creatine kinase and aspartate aminotransferase; resolution of all abnormalities occurred within 72 h (Chodorowski and Anand, 2003). In a second attempted suicide report by the same authors, a 22-year-old ingested 5 g captan and complained of nausea, weakness, numbness of the upper limbs and substernal pain (Chodorowski et al., 2004). In both cases, the mg/kg dose, using a standard 60 kg body weight, was well below the rodent LD50.

90.3.1.10  Summary Adverse skin reactions to captan and folpet due to delayed contact hypersensitivity are possible for mixers, loaders, and applicators, and may occur in low incidence. The potential for eye irritation exists, but extensive use experience suggests this problem is minimal. There is little acute risk from oral ingestion or dermal exposure to either product. The World Health Organization has classified folpet as unlikely to present an acute hazard in normal use (FAO/ WHO, 1996).

90.3.2  Subchronic Toxicity Mechanistic studies in mice have focused on changes to the duodenum and illuminate the mode of action for tumor formation. These studies confirm the irritant properties of captan and folpet, the effects of irritation to the duodenum,

and the reversibility of these effects upon cessation of treatment. Other observations in the mouse include reduced weight gain and depressed food intake at high doses; this is also seen as a general secondary effect in rat studies. Observations in rats also include hyperkeratosis and acanthosis of the esophagus and stomach, particularly for folpet. Dogs do not tolerate capsule-administered captan or folpet well; emesis is generally seen. Rabbits administered folpet dermally show marked skin irritation, and rats respond to repeated dermal application of folpet with severe skin irritation.

90.3.2.1  Mice (a) Mechanistic Studies Male CD-1 mice (four or five per group) treated with 3000-ppm captan for 1, 3, 7, 14, or 28 days showed shortened duodenal villi due to damage by captan. This effect was observable in the crypts within 3 days of treatment initiation (Tinston, 1996). Immature cells were observed at the villi tips from day 7 onward, indicating a higher turnover of cells. There was some focal gastritis and parakeratosis noted in one mouse. When captan is fed to mice at 6000 ppm for 28–90 days, villus atrophy occurs, together with a crypt hypertrophy and crypt cell hyperplasia (Tinston, 1995). CD-1 mice administered captan for 56 days as 0, 400, 800, 3000, or 6000 ppm were evaluated for proliferative changes in the duodenum (Tinston, 1995). An assessment of the duodenum was made using histopathology and a bromodeoxyuridinelabeling index to measure crypt cell proliferation. Captan induced hyperplasia of the crypt cells, an increase in the crypt cell labeling index, and an increase in the number of cells in the crypt cell population. At 3000 ppm, the villus-to-crypt height ratio decreased from 5.4 in males and 5.9 in females to 1.4 and 2.6, respectively. These observed changes are consistent with an irritant action of captan on the duodenum. The no-observed-effect limit (NOEL) for duodenal hyperplasia is 400 ppm. Male CD-1 mice treated with 5000-ppm folpet for 28 days (Waterson, 1995) were observed to have proliferative changes in the duodenum proximal to the pyloric sphincter. An inflammatory response, similar to that noted with captan, was not seen. Treatment of CD-1 mice with folpet at 0, 150, 450, and 5000 ppm for 28 days resulted in duodenum proliferative effects (Milburn, 1997). Villi length was reduced and crypt compartments were expanded, reducing the villi-to-crypt ratio. The NOEL for hyperplasia in the duodenum for males was 450 ppm (69 mg/kg per day) and the NOEL for females was 150 ppm (29 mg/kg per day). (b) Subchronic Study B6C3F1 mice were fed folpet at 0, 1000, 5000, or 10,000 ppm for 4 weeks (Rubin, 1981). There was reduced food intake and body weight gain at 5000 ppm.

Chapter | 90  Captan and Folpet

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90.3.2.2  Rats

90.3.2.3  Rabbits

Oral Wistar rats treated for 4 weeks with captan at 0, 2000, 4000, 8000, or 12,000 ppm were observed to have a dose-related decrease in body weight gain and food intake (Til and Beems, 1979). At the top two doses, there were increases in basophilia of hepatocytes in the periportal area of the liver accompanied by increases in relative liver weight. A similar finding was observed in the females at the next lower dose (4000 ppm). The relative organ-to-body kidney weights were statistically increased at all doses. Folpet Technical admixed in the diet at 0, 2000, 4000, and 8000 ppm, and fed to Fischer rats for 13 weeks produced a treatment and dose-related decrease in body weight gain and hyperkeratosis/acanthosis of the esophagus and nonglandular stomach (Sela, 1982). The NOEL for decreased weight gain was 2000 ppm in males (136 mg/kg per day) and 4000 ppm in females (291 mg/kg per day). Irritation to the esophagus and forestomach occurred at all doses and in both sexes. A variety of hematological and clinical chemistry changes were noted, but the incidence and pattern did not indicate a clear target. Folpet Technical admixed in the diet at 0, 300, 1000, 3000, or 10,000 ppm and fed to Sprague-Dawley rats for 13 weeks, followed by a 2-week recovery showed similar signs of irritation in the forestomach, primarily at 10,000 ppm, but no irritation of the esophagus (Reno et al., 1981). Following a 2-week recovery period during which the rats received a control diet, the forestomach histology returned to normal. Folpet Technical fed to B6C3F1 mice for 28 days at levels of 0, 1000, 5000, and 10,000 ppm induced a reduced body weight gain in the top two doses (Crown, 1981).

A 21-day dermal study with captan in rabbits at 0, 12.5, 110, or 1000 mg/kg per day (6 h exposure per day) resulted in a dose-related desquamation of the skin by day 21, erythema and edema at the high dose, and acanthosis and hyperkeratosis of the treated skin at all doses (Johnson, 1987).

(a) Dermal A 28-day rat study with Folpet Technical applied mineral oil at 0, 1, 10, 20, and 30 mg/kg per day to the backs of Sprague-Dawley rats 6 h per day, 5 days per week. All dose levels elicited irritation that was more severe in males than females and resulted in decreased weight gain (Dougherty, 1988). The irritation was so severe in the 30-mg/kg per day male group that application was terminated after 10 days. All adverse effects noted were related to the skin irritation induced by repeated exposures to folpet. (b) Inhalation Captan has been tested in Wistar rats by nose-only inhalation at nominal dose levels of 0.1, 0.5, 5, and 15 g/l for 13 weeks (Hext, 1989). There were deaths in males at the high dose and dose-related effects on the larynx (e.g., squamous metaplasia, squamous hyperplasia, vacuolar degeneration of squamous epithelium). The no-observedeffect concentration (NOEC) for toxic effects (other than generalized irritation) was 0.6 g/l (measured).

90.3.2.4  Dogs Beagles, two per sex per treatment group, were administered captan by capsules at 0, 30, 100, 300, 600, or 1000 mg/kg per day for 4 weeks. The results included treatment-related emesis and a dose-related decrease in food intake and body weight gain (Blair, 1987). There was an increase in relative liver weight in males at 600 and 1000 mg/kg per day and relative kidney weight in females at 1000 mg/kg per day. Some fatty changes were seen in the kidney and liver of one male at 1000 mg/kg per day. Folpet administered to two beagle dogs per sex per group at 0, 20, 60, 180, and 540 mg/kg per day for 4 weeks induced emesis (Daly, 1983). Food intake and body weights were reduced in a dose-related manner. There were, however, no histopathologic changes noted. Folpet was administered to four beagles per sex at 0, 790, 1800, and 4000 mg/kg per day for 13 weeks with gelatin capsules (Barel et al., 1985). Daily doses of 4 g/kg were well above the maximum tolerated dose and resulted in severe deterioration of the males, all of which were killed for humane reasons. One of the females at the high dose was also terminated in moribund condition. Vomiting and diarrhea were noted clinical signs and both food intake and body weight gains were reduced in a dose-related fashion. Treatment resulted in irritation of the gastric mucosa, atrophied testes, thyroid degeneration, and muscular dystrophy.

90.3.2.5  Miscellaneous Studies Rats and mice administered 3000 ppm captan had reduced immune function as measured by sheep red blood cell antibody formation after treatment for 42 days (LaFargeFrayssinet and Decloitre, 1982). Captan was reported to suppress both B- and T-cell function in mice. Wistar rats had depressed lymphocyte count and lower relative thymus weight after 3 weeks of dietary administration of captan at 1000 ppm [50 mg/kg body weight (bw) per day], initiated as weanlings (Vos and Krajnc, 1983). Additionally, rats administered pre- and postnatal captan at 750 and 2000 ppm (37.5 and 100 mg/kg bw per day) showed a decrease in secondary IgG response to tetanus toxoid in the high-dose animals (Vos and Krajnc, 1983). Evaluation of these studies by Joint Meeting on Pesticide Residues (JMPR) concluded that captan may be an immunodepressant (FAO/WHO, 1990). These dose levels, however, are high and may not be relevant to anticipated human exposure scenarios. Three consecutive intraperitoneal administrations of captan at doses up to

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15 mg/kg to Swiss Albino CD1 mice impaired CTP-catalyzed metabolism (Paolini et al., 1999). This effect was attributed to captan metabolites; however, the route of administration is not relevant for human risk assessment. In another study in which captan was administered orally at 80 mg/kg (three doses) to rats there was no affect on any of the cytochrome P450 isoenzymes (Rahden-Staron et al., 2001).

90.3.3  Chronic Toxicity One mechanistic study was conducted in mice with captan. These data are also relevant for folpet because both share a common mechanism of toxicity. Duodenal tumors in mice were seen in both chronic and oncogenic studies (discussed below). Chronic administration of folpet produced evidence of irritation to the esophagus and stomach in rats.

90.3.3.1  Mice In a mechanistic study, male CD-1 mice were administered captan at dietary doses of 0 and 6000 ppm for 3, 6, 9, 12, 18, or 20 months (Pavkov and Thomasson, 1985). Mice were examined at the end of each dosing period and, in addition, various recovery periods were evaluated (Table 90.4). One group dosed with 6000 ppm for 6 months was held for an additional 6-month recovery period; another was held for a 12-month recovery period. One group dosed with 6000 ppm for 12 months was followed after a 6- and 8-month recovery. The most characteristic pathologic findings consisted of necrotizing and proliferative changes in the nonglandular portion of the stomach (after 3 months), dilation of the small intestine, and focal epithelial hyperplasia in the proximal part of the small intestine. Focal epithelial hyperplasia was also found in controls, but the incidence was lower compared with that of the treated animals, and the localization of these foci was more caudal

Table 90.4  Captan Mechanistic Study Designa

than was the case for captan-administered mice. Diffuse hyperplasia was found only in treated mice and was not considered prerequisite for the development of focal hyperplasia. Adenomas and adenocarcinomas also developed in the small intestines of treated mice with localization in the proximal 7 cm of the small intestine; the area of localization was the same as for the focal hyperplasia. Removal of captan from the diet resulted in a significant reduction in the incidence of focal epithelial hyperplasia as compared with the incidence in concurrent lifetime-treated mice and was no greater than that in concurrent controls. The incidence of neoplasia, however, in mice in the recovery group was not significantly different from that of concurrent lifetime-treated mice, but increased in mice treated for 6 months with a recovery period of 6 months and in mice treated for 12 months with a recovery period of 6–8 months, respectively, when compared with controls. The latter increase was not found in mice treated for 6 months with a recovery period of 12 months.

90.3.3.2  Rats Rats were administered folpet at dietary concentrations of 0, 250, 1500, and 5000 ppm (Crown et al., 1989). Body weight gain and food intake were decreased at 5000 ppm. The incidence and severity of diffuse hyperkeratosis in the esophagus and nonglandular epithelium of the stomach were increased in both sexes at 5000 ppm. The stomach was also affected at 1500 ppm. The NOEL was 250 ppm (12 and 15 mg/kg per day, males and females, respectively). A folpet combined chronic toxicity/oncogenicity study from which the EPA derived a NOEL for use in chronic dietary risk assessment employed dietary dose levels of 0, 200, 800, or 3200 ppm (Cox et al., 1985). The EPA selected the 200-ppm level (9 mg/kg per day) as the NOEL for conducting chronic dietary risk assessments, based on hyperkeratosis/acanthosis and ulceration/erosion of the nonglandular stomach at 800 ppm (equivalent to 35 mg/kg per day; U.S. EPA, 1999b).

90.3.3.3  Dogs

Time (months) Dose (ppm)

3

6

9

12

18

20

0

S

S

S

S

S

S

6000

T, S

6000

T

T, S

RS

RS

6000

T

T

T

T, S

RS

6000

T

T

T

T

T, S

6000

T

T

T

T

T

RS

T, S

T  treatment; S  sacrifice; RS  recovery sacrifice. a Reproduced with permission from Pavkov and Thomasson (1985).

Dogs were treated with captan at 0, 12.5, 60, and 300 mg/kg per day for 1 year (Blair, 1988). Only the high-dose animals differed from control in increased incidence of emesis and soft stool, increased relative liver weight, and decreased total serum protein and albumin. There were two 1-year dog studies conducted with folpet: the first at 0, 10, 60, and 140/120 mg/kg per day (Daly and Knezevich, 1986) and the second at 0, 325, 650, and 1300 mg/kg per day (Waner, 1988). In the first study, the 140 mg/kg per day was reduced to 120 mg/kg per day on day 50 due to unacceptable decreases in body weight gain and food intake. A NOEL was selected for the study at 10 mg/kg per day based on lowered body weight gain and food intake

Chapter | 90  Captan and Folpet

at 60 and 120 mg/kg per day. There were no clinical signs of toxicity noted, but clinical chemistry values showed a treatment-related decrease in total plasma protein parameters and cholesterol. Organ weights were not affected by treatment, nor was there evidence of macroscopic or microscopic changes as a result of treatment. In the second study, folpet induced incidences of diarrhea, vomiting, and salivation that were associated with reduced food intake and reduced body weight gain. Testes weights were reduced in males administered 1300 mg/kg per day when compared with controls on an absolute basis, but were similar to controls when measured on a relative body weight basis. The NOEL for this study was 325 mg/kg per day, based on decreased body weight gain. The WHO acceptable daily intake (ADI) is 0.1 mg/kg per day for both captan (FAO/WHO, 1990) and folpet (FAO/WHO, 1996).

90.3.4  Developmental and   Reproductive Toxicity 90.3.4.1  Developmental Studies (a) Rats Captan administered to Sprague-Dawley CD rats at 0, 18, 90, and 450 mg/kg per day resulted in decreased maternal weight gain and decreased food consumption at the high dose (Rubin, 1987). There were no effects on postimplantation loss or fetal survival. Fetal body weight was reduced and the incidence of “small” fetuses (3.0 g) was increased at the high dose. There were no increases in incidences of treatment-related malformations. The incidence of minor skeletal variations, including the presence of a fourteenth (lumbar) rib, incomplete fusion of vertebral hemicentra fusion, and reduced ossification of the pubes was increased at 450 mg/kg per day. The NOELs for maternal and developmental toxicity in this study were 18 and 90 mg/kg per day. In another study, folpet was administered by gavage to Sprague-Dawley CD rats at 0, 150, 550, or 2000 mg/kg per day from gestation days 6–15. Maternal toxicity in the form of decreased food intake and body weight gain was observed at the mid- and high dose. Fetuses showed slight developmental retardation at 150 mg/kg per day, suggesting the NOEL was slightly below this level (Rubin, 1983). Pups from rats treated with 400 mg/kg per day from gestation days 8–10 were normal (Kennedy et al., 1968) as were rats treated with 360 mg/kg per day from gestation days 6–19 (Hoberman et al., 1983). (b) Rabbits Captan did not induce any teratogenic effects when administered to New Zealand White (NZW) rabbits at 0, 10, 40, and 160 mg/kg per day from gestation days 7–19 (Rubin and Nyska, 1987). The highest dose was toxic to both dams and fetuses. An increased incidence of minor skeletal ­variations was seen at this dose. In another study, captan

1927

was administered to New Zealand White rabbits at 0, 10, 30, or 100 mg/kg per day (Tinston, 1991). The developmental NOEL was 10 mg/kg per day based on increased postimplantation loss, reduced mean fetal weight, and increased skeletal defects in fetuses (27 presacral vertebrae) at the maternally toxic dose of 30 mg/kg per day. THPI was tested at 5, 10, or 22.5 mg/kg/day (GD 6–28) and did not induce any adverse soft tissue or skeletal effects on fetuses (Blee, 2006b). Folpet was tested in both Dutch Belted and NZW rabbits for potential developmental toxicity. The original studies (Fabro et al., 1966; Kennedy et al., 1968; McLaughlin et al., 1969) were conducted at high doses ranging from 75 to 150 mg/kg per day during gestation days 7–12, 6–16, or 6–18. These studies consistently demonstrated the absence of adverse effects. One study reported five incidences of hydrocephaly at doses that were maternally toxic (Feussner et al., 1984; one at 20 mg/kg per day and four at 60 mg/kg per day). A second study in which doses of folpet were “pulsed” failed to replicated this finding (Feussner, 1985). The most recent study employed doses of 10, 40, and 160 mg/kg per day, and confirmed the absence of folpet-induced developmental effects (Rubin, 1985b). Although a weight-of-evidence (WOE) analysis concluded folpet is not a developmental toxin, the U.S. EPA has assigned this compound an FQPA uncertainty factor of 3 based on the initial Feussner study. Although the U.S. EPA pointed to one study where hydrocephaly occurred at maternally toxic doses of 20 (one instance) and 60 mg/kg per day (three fetuses in two litters; Feussner et al., 1984), a second “pulse dose” study (Feussner, 1985) failed to replicate this finding and other developmental studies in NZW rabbits showed no evidence of teratogenicity or hydrocephaly (Fabro et al., 1966; Kennedy et al., 1968; McLaughlin et al., 1969; Rubin, 1985b). The U.S. EPA cited the Feussner study as the basis for assigning an FQPA threefold uncertainty factor (U.S. EPA, 1999b), although a WOE analysis concluded that folpet is not a selective developmental toxin (Neal, 2000). Phthalimide was tested at 5, 15, or 30 mg/kg/day (GD 6-28) and did not produce any adverse soft tissue or skeletal effects in fetuses (Blee, 2006a). (c) Mice CD-1 mice administered folpet by gavage, subcutaneously 3 (100 mg/kg per day) or by inhalation (�������������� at 624 g/m ����������� ), showed no developmental abnormalities (Courtney et al., 1983). BL6 mice treated subcutaneously and orally with folpet at 100 mg/kg per day and AKR mice treated subcutaneously at 100 mg/kg per day were judged not to have adverse developmental findings (Bionetics Research Laboratories, 1968). (d) Other Species Captan is not teratogenic in beagles when administered in the diet at 60 mg/kg per day either throughout gestation or throughout gestation plus lactation (Kennedy et al., 1975b).

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Folpet was studied in Rhesus and stump-tailed macaques as part of research on thalidomide (Vondruska, 1969). There were no malformations with folpet at doses up to 75 mg/kg per day. Thalidomide at 10 mg/kg per day produced limb defects.

90.3.4.2  Reproductive Studies In a three-generation study, COBS CD rats were treated with captan at 25, 100, 250, and 500 mg/kg per day (Schardein et al., 1982). Nonreproductive parental toxicity was seen at 100 mg/kg per day and above in the absence of reproductive effects. Pup weights were lower by 7% compared with controls at 25 mg/kg per day. A subsequent one-generation study at 0, 6, 12.5, and 25 mg/kg per day showed no effect on pup weights at 25 mg/kg per day. The NOEL selected by the EPA for use in risk assessment was 12.5 mg/kg per day, based on the weight gain depression in the three-generation study (Ghali, 1997; Schardein and Aldridge, 1982). The reference dose employed by the U.S. EPA, 0.13 mg/kg per day, is based on this NOEL and the use of a 100-fold safety factor. Folpet administered to rats at 0, 250, 1500, and 5000 ppm showed diffuse hyperkeratosis of the nonglandular epithelium of the stomach (Rubin, 1986). The NOEL for this study was 250 ppm, which averaged 24 mg/kg per day. There were no reproductive effects noted. Other two- or three-generation studies in rats with folpet also showed no adverse reproductive effects (Hardy, 1985; Kennedy, 1967).

90.3.5  Mutagenicity 90.3.5.1  Overview The issue of mutagenicity has been controversial, but with mechanistic studies in place, disparate results in vitro and in vivo have been resolved. Throughout this chapter, the rapid degradation of captan and folpet in living systems has been central to understanding their toxicology. The pattern of mutagenicity is consistent with this degradation and provides examples of how such degradation diminishes adverse effects. In vitro studies show evidence that both captan and folpet have mutagenic potential. Captan appears more potent than folpet, and both their mutagenic activity is inversely proportional to the presence of thiols in reaction vessels. Once thresholds for complete degradation are reached, in vitro activity is abolished. The large reserves of thiols present in the intact animal and the near instantaneous reaction of captan and folpet with these thiols serve to ensure complete elimination of captan and folpet before they can reach sensitive DNA targets. The net result of this rapid degradation is the absence of mutagenicity in vivo. The mechanism by which captan and folpet effect their mutagenicity is not clear; however, data suggest that thiophosgene, in addition to the parent compounds,

Hayes’ Handbook of Pesticide Toxicology

is mutagenic (Arlett et al., 1975). For in vitro systems, both frame-shift and base-pair substitutions are seen. Cytogenetic effects are seen in vitro, but positive results are not as ubiquitous as point mutations. These clastogenic effects are reduced when enzyme-enhanced rat liver extract (S-9) is present and are generally absent in vivo. The weight of evidence shows that although captan and folpet possess inherent mutagenic potential, they are not mutagenic in vivo.

90.3.5.2  Mutations (a) In Vitro Assays Table 90.5 shows results from representative in vitro assays with Salmonella typhimurium, strain TA 100, a prokaryote organism. The greater potency of captan relative to folpet is noted. Other S. typhimurium strains showed similar results. A mutation index (the ratio between induced versus spontaneous revertants) of 7.3 for captan and 6.3 for folpet was seen for strain 104 (Barrueco and de la Pena, 1988). Positive findings were generally seen with strains 98, 1535, 1537, and 1538 (Carere et al., 1978; Shiau et al., 1981; Shirasu et al., 1976), but negative findings were seen with strain 1536 (Shiau et al., 1981). Where there were marginally positive results with captan, folpet was usually negative. Strain WP2 try–hcr of Escherichia coli showed a strong response to captan and a negative response to folpet (Nagy et al., 1975) or, where both were positive, the revertants per plate were greater for captan than for folpet (Shirasu et al., 1976). Both were positive with the WP2 try– hcr– strain (Nagy et al., 1975; Shirasu et al., 1976) as well as other tests with the WP2 strain (Bridges et al., 1972; Simmon et al., 1976). Tests with Bacillus subtilis strains TK 6321 and 5211 were positive for both compounds: captan showed a greater mutagenic response than folpet (Shiau et al., 1981). Captan also induced point mutations in Aspergillus nidulans (Martinez-Rossi and Azevedo, 1987). Assays with eukaryote organisms such as Chinese Hamster cells and mouse lymphoma cells are shown in Table 90.6. These data show that captan and folpet induce mutations when measured in vitro. THPI was tested with S. typhimurium strains TA 98, 100, 1535, and 1202 as well as E. coli WP2 uvrA, and was negative (Carver, 1985). Phthalimide is inactive in S. typhimurium as well (Rideg, 1982). (b) Effect of Exogenous Thiols on Mutagenicity Assays The rat liver S-9 fraction is added to in vitro systems to simulate the metabolic capability of intact organisms. In this way, compounds that are mutagenic only after they are metabolized by cell enzyme systems are detected. Metabolism, however, appears to play no role in the expression of mutagenicity of captan or folpet; on the contrary, the addition of S-9 serves to diminish mutagenic potency. The reduced mutagenic activity following the

Chapter | 90  Captan and Folpet

1929

Table 90.5  Prokaryote Reverse Mutation: Salmonella typhimurium, Strain TA 100 Compound

Resultsa

Reference

Captan

26.7 Revertants per 108 cells/nmol, −S-9

Shiau et al. (1981)

Table 90.6  In Vitro Eukaryote Mutation Assays Assay

Compound

Results

Reference

Chinese hamster V79/Hgprt

Captan

Positive only in the absence of serum from the culture media

Arlett et al. (1975)

 @ 50 g/plate (167 nmol), S-9

Mean number of resistant colonies: 0.3 and 0.6 at 5 and 10 g/ml captan; vapor emitted from sodium bicarbonate activated captan impregnated on filter paper above the test system also induced mutations

 @ 50 g/plate (167 nmol), S-9 Folpet

7.7 Revertants per 108 cells/ nmol, −S-9

Shiau et al. (1981)

 @ 50 g/plate (167 nmol), −S-9 – @ 50 g/plate (167 nmol), S-9 Captan

26 Revertants/nmol, −S-9

Folpet

8 Revertants/nmol, −S-9

Captan

93.7 Revertants/nmol, −S-9

Folpet

15.0 Revertants/nmol, −S-9

De Flora et al. (1984)

Moriya et al. (1983)

a  S-9: Rat liver homogenate included for “metabolism” of test material. −S-9: Incubation without rat liver homogenate.

addition of S-9 is an example of the general phenomenon of thiol-related degradation of captan and folpet. The presence of sufficient thiols abolishes mutagenic activity. The addition of S-9 or rat blood prior to the addition of captan or folpet reduces or abolishes activity (Table 90.7). When cysteine is added to either captan or folpet in varying ratios, the mutagenic activity declines as the ratio increases from 0.5 to 2.5. At a ratio of 5-m cysteine to 1-m captan or folpet, mutagenic activity is abolished (Moriya et al., 1978). Glutathione provided similar protective actions when added to assay vessels in ratios of 1 or higher compared with the fungicide (Rideg, 1982). Adverse tox­ icity as well as mutagenicity in Chinese hamster V79 cells is reduced when 10% fetal calf serum is used in the standard V79/Hgprt assay (Arlett et al., 1975). (c) In Vivo Assays Armed with knowledge of how these compounds degrade, it is not surprising that mutagenicity is absent in vivo (Table 90.8).

90.3.5.3  Cytogenetic Effects The effects on chromosomes mirror the pattern of activity for mutations: clastogenic findings are seen in vitro but are generally absent in vivo.

Chinese Captan and hamster folpet CHO/Hgprt

Both compounds were positive in the absence of S-9

O’Neill et al. (1981)

Mouse lymphoma L5178Y/TK

Positive in the absence of S-9

Oberly et al. (1984)

Captan

(a) In Vitro Assays Table 90.9 lists representative in vitro cytogenetic studies with captan or folpet. The addition of S-9 to assay vessels serves to detoxify captan and folpet as it does in the mutation assays. This action is expressed by a decrease in cytotoxicity with a resulting increase in tolerated dose levels. At some point, it is expected that the threshold for detoxification is exceeded and the remaining captan or folpet can act to affect the chromosomes. These data are mixed in that some positive and some negative findings are reported. (b) In Vivo Assays Table 90.10 lists representative in vivo cytogenetic studies with captan or folpet. Micronucleus assays were conducted in CD-1 mice with both compounds and yielded negative results. Captan was administered at 40, 200, and 1000 mg/kg (Jacoby, 1985b) and folpet was administered at 10, 50, and 250 mg/kg (Jacoby, 1985a). Chlorambucil, the positive control, resulted in a significant increase in micronuclei. Captan was also negative when tested in the wing spot test in Drosophila (Osaba et al., 2002). The work by Chidiac (Chidiac, 1985; Chidiac and Goldberg, 1987) provides valuable information with regard to the genotoxicity of these compounds. The basis for this

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Table 90.7  Comparison of Captan and Folpet Mutagenicity with the Addition of Exogenous Proteina Strain and dose

Component

Captan (rev/plate)

Folpet (rev/plate)

Commentb

E. coli

None

3200

1320

Captan is more active than folpet

WP2 hcr

S-9

30

50

S-9 decreases activity

0.15  M/plate

S-9 fraction

111

60

S-9 fraction decreases activity

(45 g/plate)

20-mM cysteine Rat blood

19 32

21 19

Cysteine abolishes activity (control rev/plate 30) Blood abolishes activity

S. typhimurium

None S-9

268 6

219 8

Captan is more active than folpet S9 abolishes activity (control rev/plate 17)

TA 1535

S-9 fraction

31

35

S-9 fraction decreases activity

0.15 M/plate

20 mM

4

6

Cysteine abolishes activity

(45 g/plate)

cysteine Rat blood

6

14

Blood abolishes activity

a

Reproduced from Moriya et al. (1978).

b

Captan (0.1 ml of 1.5 M/ml) or folpet (0.2 ml of 0.75 M/ml) was incubated for 10 min at 37°C with 0.5 ml of one of the following: S-9 (containing 0.3 ml S-9/ml), S-9 fraction (S-9 mix minus cofactors), 20-mM cysteine, or rat blood diluted twice with phosphate buffer or water as control. After incubation the tester strains (0.1 ml) and agar (2 ml) were added to the test tubes and plated out. Revertants/plate were read after incubation at 37°C for 2 days.

work drew upon evidence of mutagenicity and the tumorigenic effect captan has on the mouse duodenum. It was postulated that evidence of cytogenetic damage would be seen in the duodenum after exposure to captan. This mouse bioassay was validated with known carcinogens and noncarcinogens. Nuclear aberrations (NA) consisted of micronuclei and apoptotic bodies in the crypt cells of the duodenal epithelium. Xirradiation, 1,2-dimethylhydrazine, benzo(a)pyrene (B(a)P), and N-methyl-N-nitrosourea (MNU) induced tumors in the small intestine. Each led to a dose-related increase in the incidence of NA 24 h after administration to mice. Benzo(e)pyrene and methylurea, which are noncarcinogenic structural analogues of B(a)P and MNU, did not induce NA. Cells of the duodenum were harvested and examined for the presence of NA after a variety of captan dose regimens. Captan as well as THPI consistently failed to induce NA. Captan was administered to male CD-1 mice using a number of regimens, including a single bolus dose of 4000 mg/kg, dietary dose levels of 4000 and 16,000 ppm, and five repetitive doses totaling 5000 mg/kg (Table 90.11). In all cases, including pretreatment with l-buthionin-S,R-­sulfoximine (an inhibitor of glutathione synthesis), the investigators noted an absence of the expected signs of DNA damage. Folpet was tested in a study that replicated the Chidiac experimental design (Gudi and Krsmanovic, 2001). Mice were administered five consecutive daily oral doses of folpet at 2000 mg/kg per day. Nuclear aberrations in the duodenal crypt compartment were absent in folpet-treated mice, whereas mice administered a single dose (65 mg/kg) of dimethylhydrazine showed both apoptotic cells and micronuclei in the crypts.

90.3.5.4  Dominant Lethal Assays Dominant lethal assays have generally been negative (Table 90.7). However, positive findings for both compounds have been reported (Collins, 1972a,b). In spite of these positive findings, it appears that the compounds do not induce dominant lethal effects. This conclusion is based on: (1) the absence of positive micronucleus assays; (2) the absence of adverse effects in two-generation rat reproductive studies; (3) the lack of negative dominant lethal effects in other studies [captan: Kennedy et al. (1975a), Rideg (1982), Shirasu et al. (1978), Tezuka et al. (1978); folpet: Bradfield (1980), Calandra (1971), Kennedy et al. (1975a), Rideg (1982)]; and (4) the consistency of the findings with the rapid degradation of these compounds. A less than 1-s (captan) or less than 5-s (folpet) half-life in blood argues against the possibility of parent molecules reaching the testes. Thiophosgene is considered to be more reactive than captan or folpet, but also would not reach the testes.

90.3.5.5  DNA Interaction Captan was negative for inducing unscheduled DNA synthesis (UDS) in human diploid fibroblasts in vitro (Mitchell, 1975). It was also negative for UDS in primary liver cells (Probst et al., 1981; Rocchi et al., 1980). The nature of the captan and folpet molecules imparts difficulties in conducting in vivo DNA binding studies. Two approaches, using radiolabeled test material, have sought to determine if captan covalently binds with duodenal DNA of the mouse. In both cases, the trichloromethylthio side chain was labeled because it is the chemically

Chapter | 90  Captan and Folpet

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Table 90.8  In Vivo Mutagenicity Assays

Table 90.9  In Vitro Cytogenetic Assays

Assay

Compound Results

Reference

Assay

Compound

Results

Reference

Somatic cell mutation

Captan

Negative, oral

Nguyen (1981)

Chinese hamster V79

Captan

Tezuka et al. (1980)

(Mouse spot test)

Captan Folpet

2.2% frequency after intraperitoneal dose of 15 mg/ kg Negative, oral

Imanishi et al. (1987) Moore (1985)

Positive for sister chromatid exchange and chromosomal aberrations

Captan, folpet

Positive in the absence of S-9

Ishidate et al. (1981)

Negative

Mollet and Wurgler (1974)

Chinese hamster lung fibroblasts Chinese hamster ovary (CHO)

Folpet

Positive, but higher concentrations required with S-9

Loveday (1989)

Human blood

Captan

Negative

Pilinskaya (1983)

Lymphocytes

Folpet

Negative (5 g, 2-h exposure)

Bootman et al. (1987)

Human lymphoid cell line

Captan

Positive in the absence of S-9

Sirianni and Huang (1978)

Human diploid fibroblast cell line

Captan

Negative

Sasaki et al. (1980); Tezuka et al. (1978)

Human embryonic lung and rat kangaroo cell lines

Captan

Positive

Legator (1969)

Somatic mutation and recombination (Drosophila SMART test)

Captan

Mouse heritable translocation assay

Captan

Drosophila sex-linked

Captan

Recessive lethal assay

Folpet Captan Captan Folpet Folpet

Negative

Negative

Negative Negative Weakly mutagenic Weakly mutagenic Negative

Mouse dominant

Captan

Negative

Lethal assay

Folpet Captan

Negative Negative

Folpet Captan Folpet Folpet

Negative Negative Negative Negative

Captan

Negative

Simmon et al. (1977)

Kramers and Knaap (1973) Mollet (1973) Valencia (1981) Vogel and Chandler (1974) Jorgenson et al. (1976) Kennedy et al. (1975a) Rideg (1982) Epstein et al. (1972) Simmon et al. (1977)

Table 90.10  In Vivo Cytogenetic Assays Assay

Compound

Results

Reference

Captan

Negative

Jacoby (1985b)

Folpet

Negative

Jacoby (1985a)

Captan

Negative

Tezuka et al. (1978) Fry and Fiscor (1978) Chidiac and Goldberg (1987) Esber (1983)

Rat dominant

Captan

Positive

Collins (1972a)

Micronucleus

Lethal assay

Folpet Folpet

Positive Negative

Collins (1972b) Bradfield (1980)

Chromosomal aberration

active portion of the molecule and is expected to participate in DNA binding if it occurs. The first study used 14Ccaptan (Selsky and Matheson, 1981); the second study used 35S-captan (Provan et al., 1995). When the carbon atom of the trichloromethylthio moiety is labeled, it enters the C-1 carbon pool via CO2 that is

Negative Negative (see Table 22)

Heritable translocation

Folpet

Negative

Captan

Negative

Jorgenson et al. (1976)

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formed as the molecule degrades. As such, a low level of ubiquitous labeling appears throughout the mouse. When the sulfur atom of the trichloromethylthio moiety is labeled, sulfur exchange occurs, resulting in a low level of incorporation of 35S into proteins. Histones, in turn, are associated with DNA and result in a low level of associated radioactivity (Provan et al., 1995). Investigators have concluded that covalent binding of captan to DNA has not been demonstrated (Pritchard and Lappin, 1991; Provan et al., 1992, 1993; Selsky and Matheson, 1981). An experimental design

Table 90.11  Nuclear Aberration Study with Captana Treatment

Dose levels

Results

Single bolus dose

0 and 4000 mg/kg

Negative

Single bolus dose after pretreatment with BSOb

0 and 4000 mg/kg

Negative

Dietary administration, 7 days

0, 8000, and 16,000 ppm

Negative

Five daily doses

Single 0 20 200 1000

Negative

Cumulative 0 100 1000 5000c

a

Reproduced with permission from Chidiac and Goldberg (1987). BSO: L-buthionin-S,R-sulfoximine, an inhibitor of glutathione synthesis. c Doses: Day 1, 2000 mg/kg; day 2, dosing suspended due to toxicity; days 3–5, 1000 mg/kg. b

that “proves the negative,” however, has not been achieved and the EPA holds that in vivo DNA binding has not been ruled out (Hsu and McCarroll, 1998). The polyps, adenomas, and adenocarcinomas that develop in the mouse duodenum as a result of continuous oral administration of captan or folpet arise from the crypt cell compartment. Figure 90.5 depicts a simplified anatomy of the duodenum. Should mutagenicity play a role in the development of these tumors, it must be consistent with this anatomy and the nature of the chemical reactions associated with these compounds. There are two factors that suggest mutagenicity cannot be involved in the etiology of these tumors: first, nearly all absorption takes place through the villi; second, the degradation rates of captan and folpet prevent them from reaching the crypt compartment through diffusion. Material that is absorbed through the villi enters blood or lacteal vessels and is transported away from the crypts. Crypt cells receive blood supply from arterial vessels rather than the portal system. The remaining molecules that start to diffuse down to the crypt compartment must first pass through mucus and then diffuse through approximately 16 epithelial cells before reaching the stem cells located in position T4 from the base of the crypt (Potten and Loeffler, 1990). Mutational events in cells distal to the stem cells (some of which may still be dividing) are of no import because these cells migrate up the villi and are shed within 2–4 days. The duodenal mucosal cells are rich in glutathione, having a concentration of approximately 8 mmol in CD-1 mice (Chasseaud et al., 1991); thus, the degradation of parent molecules is promoted. Whereas the half-lives of captan and folpet are very short, the exponential loss of captan virtually eliminates all molecules in short order.

Migration of cells

Mucus

> 99% Absorption through villi

Villus

<1% Absorption through crypts

Crypt Basal cell location (T4)

To portal circulation Arterial supply Figure 90.5  Schematic of duodenal villi and crypts.

Chapter | 90  Captan and Folpet

A study of DNA damage in agricultural workers using the Comet assay showed the mean tail intensity and tail moment was greater for 134 agricultural workers compared with control populations. The mean levels of THPI were elevated in these workers (0.14 g/ml) compared with controls (0.078 g/ml); however, specific pesticide exposure levels were not obtained and the thrust of this work was to demonstrate the feasibility of rapid studies of DNA damage using this technique (McCauley et al., 2008). In summary, captan and folpet are chemically active molecules that can induce mutations and cytogenetic effects if they are positioned to interact with sensitive targets. The very nature of this reactivity coupled with mammalian anatomy, however, precludes such interaction in vivo. This conclusion is supported by the weight of evidence (including chemical fate), although instances of positive results in vivo have been reported. Captan and folpet are judged not to act as mutagens or genotoxins in the intact animal.

90.3.6  Carcinogenicity 90.3.6.1  Overview Both captan and folpet induce duodenal tumors in mice when fed at high doses. Rodent bioassay data are robust: there is a treatment and dose relationship of tumor incidence in mice, but such a relationship is absent in rats. Rat studies, however, have shown increased incidences of some tumors, but these are judged to be not treatment-related (Gordon et al., 1994). While this finding was not initially embraced by the U.S. EPA (Quest et al., 1993; U.S. EPA, 1999a,b), the current classification of captan agrees with this finding (U.S. EPA, 2004) and concludes: The new cancer classification considers captan to be a potential carcinogen at prolonged high doses that cause cytotoxicity and regenerative cell hyperplasia. These high doses of captan are many orders of magnitude above those likely to be consumed in the diet, or encountered by individuals in occupational or residential settings. Therefore, captan is not likely to be a human carcinogen nor pose cancer risks of concern when used in accordance with approved product labels.

Similar analysis has been completed for folpet (Cohen, 2008) and these data are being submitted to the U.S. EPA under PRIA for cancer reclassification. Reliance on tumor incidence reported in U.S. EPA Data Evaluation Records (DERs) in the absence of weight of evidence MOA analysis can result in questionable data. For example, the U.S. EPA’s ToxRefDB notes folpet induces multisite tumors in rats and captan induces multisite tumors in mice (Martin et al., 2009).

90.3.6.2  Mouse Bioassays An early captan study combined both gavage and dietary administration (Innes et al., 1969). These investigators

1933

dosed neonatal F1 hybrid mice by gavage with 215 mg/kg per day for 3 weeks and followed by dietary administration of 560 ppm for 18 months. This study was negative. However, the dietary concentration, in hindsight, appears to be below the threshold necessary for tumor induction. The National Cancer Institute administered captan to B6C3F1 mice at 8000 and 16,000 ppm for 80 weeks (NCI, 1977). Duodenal tumors were evident at the high dose (three in 46 males; three in 48 females). There was one in 43 males at the 8000 ppm that also had a tumor. Two other studies confirmed the treatment relationship of captan and duodenal tumors (Daly and Knezevich, 1983; Wong et al., 1981). The tumor incidence is shown in Table 90.12. The NOEL for duodenal tumors in mice (based on proliferative changes in the duodenum) is 400 ppm. Captan has also been evaluated by intraperitoneal and dermal administration. A study that treated two different “strain A” mice intraperitoneally with captan (along with 64 other chemicals previously tested by the National Cancer Institute) indicated a slight increase in lung tumors in males in one strain (Maronpot et al., 1986), but the significance of these data was questioned due to lack of interlaboratory consistency and lack of correlation to the standard rodent bioassays (FAO/WHO, 1990). A dermal study using the two-stage carcinogenesis model concluded that captan was neither a complete skin carcinogen nor a promoter, although at high doses (450 mg/kg three times per week for 3 weeks, followed by croton oil factor A1 three times per week for 51 weeks) there was some evidence it may act as a weak initiator (Antony and Mehrota, 1994). Tissue damage rather than mutagenic effect might account for this finding, however, because the control, DMSO, did not replicate the irritation effects of captan. The carcinogenic effect of folpet on the mouse duodenum is similar to that of captan (Table 90.13). The first two bioassays had doses of 1000, 5000, and 10,000 ppm (Rubin, 1985a), and 1000, 5000, and 12,000 ppm (Wong et al., 1982). In both cases, there was a low incidence of tumors at 1000 ppm. This finding triggered a third study with lower doses (East, 1994). The NOEL for tumors in the third study was established at 450 ppm. Although the primary site of gastrointestinal tumors is the duodenum in mice, a low incidence of tumors is seen in the stomach with folpet. There were some tumors noted with captan, but the incidence was low, not dose-related, and not obviously treatment-related. The differential aqueous stability in acid conditions of the stomach between captan and folpet may account for this finding. Captan elicits effects in the stomach, but these are restricted to polyp formation. The blockage from the stomach to the duodenum seen in some mice that resulted from the presence of polyps and tumors located just after the pyloric sphincter was suggested as a contributing cause of stomach tumors (Nyska et al., 1990). This blockage was thought to result in an increased concentration of folpet and folpet degradates

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Table 90.12  Captan Duodenal Tumor Incidence in Micea Dose (ppm and mg/kg/day) 0

100

400

800

6000

8000

10,000

16,000

0

15

61

123

925

NC

NC

NC

Adenoma

2/91

3/83

0/93

1/87

4/84

Carcinoma

0/91

0/83

0/93

0/87

2/84

Adenocarcinoma

0/9

Reference

Males Daly and Knezevich (1983)a

3/43

5/46

NCI (1977) Wong et al. (1981)b

Duodenal neoplasms 2/74

20/73

21/72

39/75

0

100

400

800

6000

8000

10,000 16,000

0

18

70

142

1043

NC

NC

Adenoma

3/85

1/82

1/83

7/81

3/91

Carcinoma

0/85

0/83

0/83

0/81

1/91

Adenocarcinoma

0/9

NC

Females Daly and Knezevich (1983)c

0/49

3/48

NCI (1977) Wong et al. (1981)b

Duodenal neoplasms 2/72

24/78

19/76

26/76

a

NC: not calculated. Incidence reported reflects pathology reevaluation of slides (Robinson, 1993). c The total tumor incidence combines both benign and malignant tumors. b

Table 90.13  Folpet Duodenal Tumor (Adenoma/Carcinoma) Incidence in Mice Dose (ppm and mg/kg/day) 0 0 Males

47

1000 93

b

1350 151

5000a 502

1/87

2/61

8/67

0

150 16

0/42

450 51

1000 96

b

1350 154

5000 515

a

10,000

b

0/88

1/63

7/67

12,000 1284

19/52

Rubin (1985a) 38/73

1/50

Dose levels for the Rubin study were 5000 and 10,000 ppm for the first 21 weeks and then adjusted down to 3500 and 7000 ppm. Wong et al. (1982).

b

Wong et al. (1982) East (1994)

a

10/52

0/49

Reference Rubin (1985a)

0/44

2/52

0/49

25/52 38/71

1/51

0/96

12,000 1282

17/52

0/48

10,000a

b

4/52

0

a

16

450

0/52

0/89

Females

150

Wong et al. (1982) East (1994)

Chapter | 90  Captan and Folpet

1935

Table 90.14  Rat Bioassays with Captan and Folpet Findingsa

Reference

Osborne-Mendel 0, 2525, 6060 ppm (TWA)b mg/kg/day not calculated

Negative

NCI (1977)

Wistar (Cpb:WU) 0, 125, 500, 2000 ppm 0, 6.25, 24, 98 mg/kg/day

Negative (uterus)

Til et al. (1983)

Charles River CD1 0, 500, 2000, 5000 ppm 0, 25, 100, 250 mg/kg/ day

Negative (kidney)

Goldenthal et al. (1982)

0, 1000, 5000, 10,000 ppm mg/kg/day not calculated

Negative

Hazleton (1956)

Fischer (chronic toxicity study) 0, 250, 1500, 5000 ppm 0, 12.5, 75, 250 mg/kg/day

Negative

Crown et al. (1989)

CD 0, 200, 800, 3200 ppm 0, 10, 40, 160 mg/kg/day

Negative (thyroid, testes)

Cox et al. (1985)

Fischer 0, 500, 1000, 2000 ppm 0, 25, 50, 100 mg/kg/day

Negative (mammary glands, thyroid, lymphoma)

Crown et al. (1985)

Test material Experimental design Captan

Folpet

a The U.S. EPA notes the incidence of tumors (in tissues) “associated” with treatment in organs listed in support of B2 cancer classification for both captan and folpet (Quest et al., 1993). Weight of evidence analysis shows captan is not a rat carcinogen (Foster and Elliott, 2000) nor is folpet (Study Director conclusions). b TWA: time-weighted average.

in the stomach. Stomach tumors were evident, however, where no blockage was apparent and thus argue against this hypothesis (East, 1994).

1985, 1989). The U.S. EPA now concurs with these findings for captan (U.S. EPA, 2004) and is reevaluating the folpet data.

90.3.6.3  Rat Bioassays

90.3.6.4  Comparison of Rat and Mouse Response to Folpet

In contrast to mice, there is no consistent tumor response across studies with rats (Table 90.14). Evaluation of captan tumor incidence data for kidney and uterine tumors using appropriate statistics and proper tumor grouping shows no treatment effect (Foster and Elliott, 2000; Gordon et al., 1994). It is unlikely the kidney tumors are related to treatment with captan because there is no increase in malignant tumors (carcinomas), there is a small increase (in a single animal) in benign tumors (adenomas) only, there is no statistically significant increase or trend in kidney adenomas, and the finding of kidney adenomas is seen in one out of four rat bioassays with captan and one out of seven bioassays with both captan and folpet. It is unlikely the uterine sarcoma tumors are related to treatment with captan because there is no statistical significance when tumors and polyps are considered together, a consideration dictated by the etiology of uterine sarcomas (Leininger and Jokinen, 1990). The four tumors noted comprise three different cell types and this finding was not consistent with the other bioassay results. In evaluating this study, the JMPR found “no other effects” in addition to depression of food intake and body weight gain at 2000 ppm and a slight increase in relative liver weight in males (FAO/WHO, 1990). With folpet, the study director concluded that the incidence for mammary glands and thyroid tumors were not related to treatment (Crown et al.,

A stark difference between mice and rats exists when comparing their tumor response to captan and folpet. All strains of mice tested show a treatment and dose-related incidence of duodenal tumors. All strains of rats show neither duodenal tumors nor proliferative changes. This suggests that the physiology of the mouse and rat differ in specific toxicokinetic and/or toxicodynamic ways that account for this difference. A series of comparative studies in the CD-1 mouse and Sprague-Dawley rat were conducted with folpet at 50 and 5000 ppm in an attempt to uncover the reason or reasons for this difference (Chasseaud et al., 1991). Areas investigated included respective milligrams per ­ kilogram per day doses, transit times through the gut, glutathione changes with dosing, effects on enzymes, pH changes, GSH levels, and binding of 14C-folpet to intestinal components. There were a variety of quantitative differences between rats and mice, suggesting that differential amounts of available glutathione may influence the tumor response.

90.3.6.5  Relevance of Mutagenicity to Mouse Tumors and Human Risk Assessment The presence or absence of a mutagenic component in the etiology of mouse duodenal tumors determines the paradigm used to assess risk to humans. Currently, the U.S. EPA

1936

acknowledges that mutagenesis is not involved in tumor induction by captan; thus, a threshold-based margin of exposure risk assessment is appropriate. As folpet is still officially classified as a B2 carcinogen (pending ­ ongoing PRIA review), the linear multistage model (Pitot and Dragan, 1996) can be used to describe risk.

90.3.6.6  Mode of Action Leading to Duodenal Tumors in the Mouse There is a preponderance of data that point unambiguously to a proliferation-based nongenotoxic mode of action for captan and folpet. This mode of action is consistent with the chemical and physical properties of captan and folpet, and the effect these compounds produce with high dietary exposure in the mouse. A key component of this mode of action is that it is threshold-based; that is, at dietary doses below the threshold, tumors will not develop. The physical and chemical properties include the instability of captan and folpet in aqueous solution at physiological pH, the reaction of captan and folpet with thiols, the generation of thiophosgene from both hydrolysis and thiol interactions, the transient nature of thiophosgene due to its chemical reactivity, and the comparatively low toxicity of THPI and phthalimide. The mode of action for mouse duodenal tumors must be consistent with these properties and effects. This MOA must also be supported by generally accepted principles of carcinogenicity. Mouse duodenal tumors develop with oral administration above a threshold if maintained for at least 6 months (Pavkov and Thomasson, 1985). Histopathological analysis shows that tumors arise from the crypt compartment and show a continuum from hyperplasia to polyps, adenomas, and adenocarcinomas (Tinston, 1995, 1996). Histologic and proliferation studies have characterized the changes to the duodenum with exposure to captan and show two sequential events (Allen, 1994; Foster, 1994). First, epithelial cells that comprise the villi are damaged by exposure to captan and sloughed off into the intestinal lumen at an increased rate. The villi height is shortened. Second, basal cells in the crypt compartment that normally divide at a rate commensurate with the normal loss of villi cells from the tips of the villi increase their rate of proliferation to a hyperphysiologic state. Crypt depth is subsequently increased and the villi-to-crypt ratio (measured by their respective sizes) decreases. A small number of transformed cells exist in the duode­ num as evidenced by a low incidence of duodenal tumors in bioassay control (Tables 90.12 and 90.13) and historic­al control mice (Bomhard and Mohr, 1989; Chandra and Frith, 1992; Lang, 1995; Maita et al., 1988; Ward et al., 1979). It is postulated that these transformed cells are subject to proliferative pressure and, as a result of this continued pressure for at least 6 months, progress to tumors. The basis for this postulation is the body of data that

Hayes’ Handbook of Pesticide Toxicology

show abnormally high cell proliferation, which is not carcinogenic per se, but does play a role in tumor development (Butterworth et al., 1992; Ledda-Columbano et al., 1989; Pitot et al., 1991). The role of proliferation in thyroid tumors is well established (Chhabra et al., 1992) and the influence of proliferation on initiated liver cells is also known (Solt et al., 1977). Classically, the two-stage carcinogenesis model in the skin points to the importance of sustained proliferation in the promotion of initiated cells to tumors (Berenblum and Armuth, 1977). In addition to promoting the clonal expansion of nascent tumor cells in situ, abnormally high proliferation may increase fixation and expression of premutagenic DNA lesions, increase the number of spontaneously initiated cells during replication, perturb checkpoints in the cell cycle leading to mutagenic events, and increase the number of spontaneously initiated cells by blocking cell death/ elimination (Ledda-Columbano et al., 1989). Thus, there are two avenues for duodenal tumors to develop: promotion of nascent tumor cells and initiation of normal basal cells through disruptions in normal DNA replication. The progression to tumors under this mode of action is depicted schematically in Figure 90.6. A genetic component is neither required nor plausible. Thresholds have been established for the initial cellular response to captan or folpet administration: villi damage and crypt cell hyperplasia. The NOELs for captan and folpet are similar: 400 ppm (60 mg/kg per day) for captan and 450 (69 mg/kg per day; males) or 150 ppm (29 mg/kg per day; females) for folpet. Administration of captan or folpet below these thresholds will not lead to tumors, because the basis for tumor progression (hyperphysiologic cell division rate) is absent. This mode of action requires that the appropriate paradigm for assessing carcinogenic risk in humans is margin of exposure not linear low-dose * extrapolation ( q1 ).

90.3.6.7  Epidemiology In a limited retrospective cohort mortality study, 138 workers in a captan manufacturing plant who were employed for a minimum of 3 months during a 23-year period beginning in 1954 were followed for 30 years (Palshaw, 1980). These workers were exposed to captan at estimated air concentrations ranging from 0.83 mg/m3 for THPI operators to 1.54 mg/m3 for captan operators. Other workers had little or no exposure (originally ranked as 0, 1, 2, or 3 for none, low, moderate, or high, respectively). These data showed there were no increased deaths that resulted from captan exposure. An analysis of epidemiology data from the Agricultural Health Study concluded that observed cancer cases and follow-up times of 9.14 years, “though limited by low numbers,” did not provide evidence of increased cancer risk due to use of captan (Greenburg et al., 2008).

Chapter | 90  Captan and Folpet

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Captan/folpet

Shortened villi Villi cell loss

Enlarged crypts

Crypt cell proliferation Hyperplastic crypts Dose > 50 mg/kg/day

Normal duodenum

Removal of captan/folpet Rapid recovery

Normal duodenum Captan/folpet Continued irritation Adenoma

Adenocarcinoma

Proliferative pressure on spontaneously-transformed cells in situ. Figure 90.6  Mode of action for captan and folpet in the mouse duodenum.

90.3.6.8  Summary Captan and folpet at sufficiently high doses act locally on the duodenal mucosa and result in damage to villi. Epithelial cells of the villi are lost and the homeostatic feedback mechanism increases cell proliferation in an attempt to make up this loss. Transformed cells that reside in the crypt compartment are sensitive to this proliferative pressure and are promoted to frank tumors (Figure 90.6). This mode of action has no mutagenic component and has a clear threshold for the first event that leads to tumors: increased ­proliferation/ hyperplasia of the duodenal crypt compartment.

90.4  Common mechanism of toxicity 90.4.1  Captan and Folpet Captan and folpet show obvious similarities in structure and effects. The Food Quality Protection Act (U.S. Congress, 1996) formally recognized the existence of such similarities and mandated that the U.S. EPA consider common mechanisms of toxicity when conducting risk assessments. The U.S. EPA issued guidance on how to determine the presence of a common mechanism for two or more pesticides (U.S. EPA, 1999c). Their criteria include structure, adverse effects, and mode of action. Captan and folpet

share sufficient common characteristics to conclude that they have a common mechanism of toxicity (Bernard and Gordon, 2000). This finding is specific to the key toxicological endpoint, duodenal tumors in mice, but may apply as well to other nonspecific endpoints. The finding that a common mechanism of toxicity exists for captan and folpet is supported by the following determinations: 1. Structural similarity. The active side chains, —SCCl3, are identical. 2. Site of action. Toxicity is expressed at the site of contact for both chemicals (that is, they are local irritants as opposed to systemic toxicants). 3. Reactivity with thiols. Both react with thiols to produce similar degradates. Differences in rates of reaction are attributable to the physical/chemical properties of the two compounds and do not serve to diminish their commonality. 4. Mechanism of pesticidal action. Toxicity to fungi is mediated through reactions with both soluble and insoluble thiols in fungal conidia. These same reactions account for expression of the common toxic endpoint in mammals. 5. Common toxic endpoint. Gastrointestinal tumors in mice that generally are specific to the duodenum. 6. Mode of action. Both captan and folpet express their common toxic endpoint through a nongenotoxic compensatory proliferation mechanism.

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7. Specificity of action. For both materials, the majority of tumors appear in the duodenum, but with folpet some tumors are noted in the stomach. The hydrolytic rate of folpet is approximately eight times faster than that of captan at pH 5 and may promote the presence of active metabolites in the acid environment of the stomach. Tumors are restricted to the mouse; rats are refractory. 8. Other toxic endpoints. Both captan and folpet show a similar pattern of toxicity for mutagenicity and skin sensitization. Both compounds show nonspecific secondary endpoints such as developmental toxicity manifested as decreased fetal weights and ossification defects at maternally toxic doses. Finding a common mechanism of toxicity for captan and folpet will influence the way the U.S. EPA regulates these two fungicides under FQPA. It also will afford toxicologists an opportunity to integrate data from the individual compounds to generate a more robust database, which is particularly valuable for evaluation of noncarcinogenicity in rats and elucidation of the mode of action of carcinogenicity in mice.

90.4.2  Captafol Captafol (CAS 2939-80-2; Figure 90.7) differs from captan and folpet in a number of areas. The side chain differs in structure as well as chemical activity. The two-carbon tetrachloro moiety of captafol is able to produce an episulfonium ion that can act as a systemic alkylating agent (Figure 90.8). This ion, absent with captan and folpet, is

O

– –

Cl Cl

– –

N— S– C –C–H Cl Cl

O Captafol

–

–

– –

CH3

F

–

CH3–

–

–

CH3

Cl O O N— S—N— S– C –Cl –

–

Cl O O N—S—N— S– C –Cl –

CH3–

F

CH3 Tolylfluanid

Dichlofluanid

Figure 90.7  Captafol, dichlofluanid, and tolylfluanid.

N— S +–C–CCl2

–

Captafol

N— S– C –C–H – –

O

Cl Cl

Cl –

Cl Cl

– –

– – – –

Cl Cl

N—S–C–C–H

–

O

CHCl

Cl Cl

Episulfonium ion

Figure 90.8  Episulfonium ion formation by captafol.

able to enter the systemic circulation and may be carcinogenic (Williams, 1992). The spectrum of tumors in rodent bioassays is broad and affects both mice (Ito et al., 1984) and rats (Nyska et al., 1989; Quest et al., 1993), whereas the tumor spectrum of captan and folpet is narrow, focusing on the mouse duodenum (Gordon et al., 1994). Mutagenic results in some assays show a differing pattern of activity. For example, when tested in S. typhimurium, TA 102 and TA 104, captan was negative in strain TA 102 and positive in strain TA 104, whereas captafol was negative for TA 104 and positive for TA 102 (Barrueco and de la Pena, 1988). In S. typhimurium strains TA 100, TA 98, TA 1535, TA 1537, and TA 1538 as well as E. coli strain WP2 hcr, captan and folpet were positive in all systems, whereas captafol was positive only in WP2 hcr and was “doubtful” in TA 100 (Moriya et al., 1983). Two results follow from the finding that captafol does not share a common mechanism of toxicity with captan and folpet. First, under FQPA, residues will not be combined for a cumulative risk assessment. Second, the “structural similarity” (Quest et al., 1993) of captafol should not be referenced when evaluating the carcinogenicity of captan or folpet. The first point is moot, because captafol is not registered in the United States; the second point avoids confounding comparisons.

90.4.3  Dichlofluanid and Tolylfluanid Dichlofluanid (CAS 1085-98-9) and tolylfluanid (CAS 73127-1) do not share a common mechanism of toxicity with captan or folpet with regard to mouse duodenal tumors, principally because they do not induce these tumors. Both compounds have a fluorine atom substituted for one of the three chlorine atoms on the trichloromethylthio moiety (Figure 90.7). They differ from one another by the addition of a methyl group on the benzene ring. Like captan and folpet, these compounds react with sulfhydryl groups (Schuphan et al., 1981). The monofluorodichloromethylthio moiety conveys more chemical reactivity to the parent as measured by the reaction rate with 4-­nitrothiophenol compared with the trichloromethylthio moiety. The reaction rate of dichlofluanid is over twice that of captan and folpet, but the trichloro dichlofluanid analogue is less reactive than either captan or folpet. Dichlofluanid and its bis(fluorodich loromethyl) disulfide degradate were reported to be negative for mutagenicity in S. typhimurium TA100, whereas the bis-(trichloromethyl) disulfide from captan and folpet was positive (Schuphan et al., 1981). The presence of the fluorine atom apparently lessens the mutagenicity of these compounds. Thiophosgene and its monofluorine analogue are postulated to be degradates of dichlofluanid. Either compound reacts with cysteine to form TTCA in a similar way as it is formed with captan or folpet. These compounds have been reviewed by the Joint Meeting of the FAO Panel of Experts on Pesticide Residues

Chapter | 90  Captan and Folpet

in Food and the Environment and the WHO Expert Group on Pesticide Residues (FAO/WHO, 1984, 1989). Dichlofluanid was negative for carcinogenicity when tested in mice at 5000 ppm. The levels that cause no toxicological effect in rats and dogs are 500 (30 mg/kg bw per day) and 1000 ppm (25 mg/kg bw per day), respectively. For tolylfluanid, the levels that cause no toxicological effect in rats and dogs are 300 ppm (15 mg/kg bw per day) and 12.5 mg/kg bw per day, respectively. The absence of duodenal tumors in mice suggests that the ability to induce these tumors is not a general property of the chloroalkylthio fungicides.

90.5  Human risk assessment 90.5.1  Cancer In contrast to the relatively high background duodenal tumor incidence seen in mice, the incidence in humans (Parkin et al., 1992) and rats (Goodman et al., 1979; Maekawa et al., 1983; Maita et al., 1987; McMartin et al., 1992) is low. This suggests that humans are closer to rats; that is, humans are refractory to tumors with captan or folpet because the number of transformed cells in situ is low. Nonetheless, prudence dictates that humans be considered similarly to the mouse for risk assessment purposes. The no effect levels for duodenal crypt cell proliferation, the prerequisite for tumor formation, are 400 ppm for captan and 150–450 ppm for folpet. The approximate equivalent doses are 30–60 mg/kg per day. For this assessment we used a NOEL of 50 mg/kg per day. Humans are exposed to captan and folpet predominantly by two routes: oral and dermal. Exposure via the oral route occurs through consumption of food that contains residues; exposure via the dermal route occurs through the use of products that contain these fungicides. Exposure from food is low and there are no contributions from water. Milk, which is both aqueous-based and metabolically-produced, was shown to have no captan or degradates present. A national milk survey for captan that was conducted over the course of 1 year and analyzed 224 samples from a statistically derived paradigm across four regions of the United States (North East, North Central, West, and South) found no detectable levels (LOQ  0.005 ppm) of captan, THPI, 3-OH-THPI, or 5-OH-THPI (Slesinski and Wilson, 1992). Exposure to oral residues only is considered relevant for human cancer risk assessment. Dermal exposure is not relevant for human cancer risk assessment, because dermal contact does not result in systemic exposure and captan has been found not to be a skin carcinogen (Antony and Mehrota, 1994). For both captan and folpet, the EPA has calculated the estimated exposure for cancer risk purposes as 0.00005 mg/kg per day (U.S. EPA, 1999a,b). The MOE for each of these fungicides, based on a NOEL for duodenal crypt cell proliferation of 50 mg/kg per day is 1,000,000. These MOEs suggest virtually no

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risk of cancer to persons who consume produce treated with either captan or folpet. It is unlikely that both compounds would be present on the same commodity at the same time because the uses of captan in the United States do not overlap those of folpet. Additionally, normal agronomic practice usually relies on one or the other, not both. Nonetheless, if the expected residues are combined for a cumulative risk assessment, the MOE is still satisfactory. This analysis shows that humans are not at risk for duodenal tumors from these fungicides, a position embraced by EPA for captan (U.S. EPA, 2004). This assessment is particularly relevant for re-entry workers such as strawberry harvesters who might be exposed dermally to captan residues. There are, however, instances of reported associations of captan with human cancer. In an epidemiology study that examined pesticide use and breast cancer among 30,454 farmers’ wives, there was a “possible” increase in risk associated with captan but the authors cautioned that further follow-up of this cohort was necessary to clarify the relationship (Engel et al., 2005). In a multicenter case– control study, it was reported that captan was associated with incidence of non-Hodgkin’s lymphoma (McDuffie et al., 2001). The odds ratio (OR) reported for individuals handling captan “greater than two days/year” was 2.80 (95% CI: 1.13–6.90). These associations are not discussed in light of biological feasibility. Since captan is not present systemically and since its degradate, THPI, has shown no evidence of carcinogenicity in rodent bioassays (through testing of its parent), other than local duodenal tumors, biologic feasibility for the association of captan and human cancer remains speculative.

90.5.2  Noncancer For noncancer risks, captan and folpet present an interesting challenge for risk assessors. The transient nature of these molecules coupled with their inherent low toxicity make it difficult to assign meaningful endpoints. Noncancer endpoint risk characterization requires the selection of relevant endpoints for nondietary and dietary exposure, and that NOELs be determined for both acute and chronic exposure. Nondietary exposure, in turn, comprises dermal exposure (including eye exposure) and inhalation. Three nondietary hazards associated with captan and folpet that are relevant to human safety are skin sensitization, eye irritation, and lung irritation. Only one of these, skin sensitization, appears to affect persons who come in contact with these materials. The incidence of sensitization reactions is below 10% in trials with captan and well below this incidence in actual use (Krieger, personal communication). Systemic toxicity from dermal exposure is lacking due to the labile nature of these molecules in the blood. Skin irritation from single-instance contact with captan or folpet is not expected. Repetitive dermal exposure, however, might induce progressive skin irritation, although it is not

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evident with the limited number of people who repeatedly use a shampoo containing 7% captan (Guo, 2001). Inhalation is a potential avenue for adverse effects, although the absence of adverse reports suggests that this is not an issue. The AIHGH has assigned a threshold limit value of 5 mg/m3 (5 g/l) for captan (ACGIH, 1998) and the same value has been suggested as appropriate for folpet (Seifried, 1996). In vitro studies with human bronchial epithelial cells (16HBE140-) note adverse effects due purportedly to lipid peroxidation and the generation of reactive oxygen species (ROS) at levels between 2.89 and 5.11 g/cm2 for two folpet commercial products (Canal-Raffin et al., 2007, 2008). These researchers note that folpet, used extensively in French vineyards, is relatively persistent when introduced to cell cultures. Average airborne residues in rural and urban settings associated with folpet usage in vineyards are reported to be 1.2 and 0.01 g/m3, respectively (Chretien, 2004b), which, when compared with a NOAEL of 0.0006 g/m3 (0.6 g/l) from a 90-day inhalation study with captan, provide high margins of exposure (folpet’s NOAEL is expected to be similar to that of captan). Higher average airborne residues are noted within vineyards during spray operations (40 g/m3, Chretien, 2004a), but appropriate respiratory protection is indicated for these operations. For acute dietary risk, the U.S. EPA has used the NOEL from developmental studies for both captan (U.S. EPA, 1995a) and folpet (Levy et al., 1997). For captan, this is 10 mg/kg per day, based on effects at 30 mg/kg per day (a maternally toxic dose) in a rabbit study. For folpet, this is 10 mg/kg per day, based on effects at 20 mg/kg per day in a rabbit study. This “default” selection is not ideal because the NOEL is based on multiple doses, it is based on effects on the fetus and not the individual, and it is specific for a subgroup (women of childbearing age) that comprises only part of the general population. A meaningful acute dietary risk assessment is dependent on an appropriately designed single-exposure oral toxicity study; such data are not currently at hand. Captan acute dietary exposure at the 99th percentile is estimated for the general U.S. population at 0.009512 mg/kg per day (Kidwell and Watters, 1999); the exposure for folpet is estimated at 0.00046 mg/kg per day (Petersen, 1997). The EPA estimates these acute dietary exposures at 0.036 mg/kg per day for captan at the 99.9th percentile (U.S. EPA, 1999a) and at 0.001532 mg/kg per day for folpet at the 99th percentile (U.S. EPA, 1999b). For chronic U.S. EPA estimates, the dietary exposure for the general population in the United States is at 0.000664 mg/kg per day for captan and at 0.000053 mg/kg per day for folpet (U.S. EPA, 1999a,b). For chronic dietary risk assessment, the captan NOEL of 12.5 mg/kg per day and the folpet NOEL of 9 mg/kg per day are used. Margins of exposure (NOEL ÷ exposure) for captan are 18,825 and for folpet are 169,811. The WHO ADI is 0.1 mg/kg per day for both captan (FAO/WHO, 1990) and folpet (FAO/WHO, 1996). This is approximately equal to the U.S. EPA’s cPAD for

captan, 0.13 mg/kg per day, and the EPA’s PAD (without the threefold FQPA safety factor) for folpet, 0.09 mg/kg per day. The JMPR reevaluated captan and folpet in 2007 and noted the ADI for both compounds at 0–0.1 mg/kg bw and the acute reference dose for captan at 0.3 mg/kg bw and for folpet at 0.2 mg/kg bw, both applicable only to women of childbearing age (FAO/WHO, 2007).

Conclusion Captan and folpet are structurally similar molecules that act through a common mechanism with regard to their ability to induce duodenal tumors in mice. The mode of action has been elucidated for these tumors and is dependent on irritation to and cell loss from the intestinal villi, followed by a compensatory increase in proliferation within the crypt compartment. This proliferative pressure, with time, promotes transformed cells that are normally resident in situ. The mode of action is not dependent on a mutagenic component nor are mutations within basal cells of the crypts a plausible occurrence. Captan and folpet are, however, mutagenic when tested in a variety of in vitro systems, and this observation has challenged investigators to solve the paradox that exists between in vitro and in vivo test results. The solution to this question is the finding that these compounds degrade extremely rapidly when thiols are present. In human blood, captan’s t1/2 is less than 1 s and folpet’s is less than 5 s. Thiophosgene, the reactive degradate that is formed from the trichloromethylthio side chain, reacts not only with thiols, but with other functional groups and its t1/2 is less than 0.6 s. The import of this rapid degradation is that systemic exposure to captan, folpet, or their common degradate, thiophosgene, is absent. This, along with the low estimated dermal absorption rate of 0.5% per hour, assures that adverse systemic risk in agricultural workers is absent. Local effects due to irritation, however, may occur. These include eye and skin irritation, skin sensitization, and irritation of the airways. Oral exposure at sufficient doses will irritate the mucus membranes of the gastrointestinal tract. Systemic effects noted in laboratory studies such as depressed weight gain or delayed development of fetuses and pups are secondary effects that result from the primary irritation of the gastrointestinal tract. Thus, captan and folpet, when used in the agricultural setting are characterized as follows: They have low acute toxicity. They are not carcinogenic, mutagenic, or teratogenic. l They are neither selective developmental toxins nor are they reproductive toxins. l l

Relevant hazards are the following: Irritation of mucus membranes Sensitization after repeated exposure

l l

Chapter | 90  Captan and Folpet

Irritation of the skin after repeated exposures (specifically for folpet) l Irritation of the airways l

These products have been in use for over 55 years and experience shows that eye irritation and sensitization reactions, particularly with re-entry operations, are not problematic. In addition, a limited survey of persons using a 7% captan-based shampoo indicates that repeated use does not cause skin irritation or skin sensitization reactions. Captan and folpet remain valuable fungicides as the risks are low and benefits high.

Acknowledgments Reports with MRID numbers are available from U.S. Environmental Protection Agency, Freedom of Information Office, Ariel Rios Building, 1200 Pennsylvania Avenue, N.W., Washington, DC 20460.

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Rubin, Y. (1986). Folpan: Two-Generation Reproduction Study in the Rat. Report MAK/052/FOL, Life Science Research Israel, Ltd., Ness Ziona, Israel (MRID 40135901). Rubin, Y. (1987). Captan: Teratology Study in the Rat. Report MAK/097/ CAP, Life Science Research Israel, Ltd., Ness Ziona, Israel. Rubin, Y., and Nyska, A. (1987). Captan: Teratology study in the rabbit. Report MAK/099/CAP, Life Science Research Israel Ltd., Ness Ziona, Israel. Ruzo, L. O., and Ewing, A. D. (1988). Hydrolysis of [14C]-Folpet. Report PTRL 124, PTRL West, Inc., Richmond, CA (MRID 40818801). Sasaki, M., Sugimura, K., Yoshida, M., and Abe, A. (1980). Cytogenetic effects of 60 chemicals on cultured human and Chinese hamster cells. Sensoshuktai 20, 574–584. Sauer, C., and Seaman, L. R. (1980). Primary Eye Irritation Study in Rabbits. Test Article: Merpan 50 WP. Report 0414D, Cosmopolitan Safety Evaluation, Inc., Somerville, NJ (MRID 00045688). Sauerhoff, M. W., DeBaun, J., Miaullis, J. B., and Freudenthal, R. I. (1982). The influence of captan on levels of reduced sulfhydryls in the duodenum of mice. Toxicologist 2(1), 425. Saunders, D. S., and Harper, C. (1994). Pesticides. In “Principles and Methods of Toxicology,” (A. W. Hayes, ed.), pp. 389–415. Raven Press, New York. Schardein, J., and Aldridge, D. (1982). One Generation Reproduction Study in Rats with Captan. Report 153–190; T-10486, International Research and Development Corporation (MRID 00120315). Schardein, J., Schwartz, C., and Thorstenson, J. (1982). Three Generation Reproduction Study in Rats. Report 153–096, International Research and Development Corporation (MRID 00125293). Schuphan, I., Westphal, D., Haque, A., and Ebing, W. (1981). Biological and chemical behavior of perhalogen methylmercapto fungicides: Metabolism and in vitro reactions of dichlofluanid in comparison with captan. Am. Chem. Soc. Symp. Ser. 158, 85–96. Seifried, H. (1996). “Folpet: Threshold Limit Value,” Industrial Hygiene Consultant, Bethesda, MD. Sela, J. (1982). Folpan Toxicity in Dietary Administration to Rats for 13 Weeks. Report MAK/021/FOL, Life Science Research, Stock, England. Selsky, C. A., and Matheson, D. W. (1981). The Association of Captan with Mouse and Rat Deoxyribonucelic Acid. Report T-10435, Environmental Health Center, Richmond, CA (MRID 00093759). Shah, P. V., Fisher, H. L., Sumler, M. R., Monroe, R. J., Chernoff, N., and Hall, L. L. (1987). Comparison of the penetration of 14 pesticides through the skin of young and adult rats. J. Toxicol. Environ. Health 21. Sharma, S. (1978). Thiophosgene in Organic Synthesis. Int. J. Methods Synth. Organic Chem. 803–820. Sharma, S. (1986). The Chemistry of Thiophosgene (A. Senning, ed.), Sulfur Reports, Vol. 5. Harwood, London. Shiau, S. Y., Huff, R. A., and Falkner, I. C. (1981). Pesticide mutagenicity in Bacillus subtilis and Salmonella typhimurium detectors. J. Agric. Food Chem. 29, 268–271. Shirasu, Y. (1978). Cytogenic and dominant-lethal studies on captan. Mutat. Res. 54, 227–228. Shirasu, Y., Moriya, M., Kato, K., Furuhashi, A., and Kada, T. (1976). Mutagenicity screening of pesticides in the microbial system. Mutat. Res. 40, 19–30. Siegel, M. R. (1971a). Reaction of the fungicide folpet (N-trichloromethylthio phthalimide) with a thiol protein. Pestic. Biochem. Physiol. 1, 225–233.

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Siegel, M. R. (1971b). Reactions of the fungicide folpet (N-trichloromethylthio) phthalimide with a nonthiol protein. Pestic. Biochem. Physiol. 1, 234–240. Simmon, V. F., Poole, D. C., and Newell, G. W. (1976). In vitro mutagenesis investigations of twenty pesticides. Toxicol. Appl. Pharmacol. 37(1), 109. Simmon, V., Mitchell, A., and Jorgenson, T. (1977). “Evaluation of Selected Pesticides as Chemical Mutagens: In vitro and in vivo Studies,” Health Effects Research Series, EPA-600/1–77–028 (MRID 132582). Health Effects Research Laboratory, Stanford Research Institute, Research Triangle Park, NC. Sirianni, S. R., and Huang, C. C. (1978). Effect of the fungicide folpet on growth and chromosomes of human lymphoid cell lines. Can. J. Genet. Cytol. 20, 193–197. Slesinski, R. S., and Wilson, A. E. (1992). National Milk Survey. Report Captan 91–01, Technical Assessment Systems, Inc., Washington, DC (MRID 42458801). Solt, D. B., Medline, A., and Farber, E. (1977). Rapid emergence of carcinogeninduced hyperplastic lesions in a new model for the sequential analysis of liver carcinogenesis. Am. J. Pathol. 88, 595–618. Tezuka, H., Teramoto, S., Kaneda, M., Henmi, R., Murakami, N., and Shirasu, Y. (1978). Cytogenic and dominant lethal studies on captan. Mutation Res. 57, 201–207. Tezuka, H., Ando, N., Suzuki, R., Terahata, M., Moriya, M., and Shirasu, Y. (1980). Sister-chromatid exchanges and chromosomal aberrations in cultured Chinese hamster cells treated with pesticides positive in microbial reversion assays. Mutat. Res. 78, 177–191. Thoa, N. B., and Redden, J. C. (1995). “The HED Chapter of the Reregistration Eligibility Decision Document (RED) for Captan,” Risk Characterization and Analysis Branch Health Effects Division, U.S. Environmental Protection Agency, Washington, DC. Thongsinthusak, T., Haskell, D., and Ross, J. H. (1999). Dermal Absorption of Propargite, Bensulfuron-methyl, Captan, and Maneb in Rats. Technical Report HS-1792, Worker Health and Safety Branch, California Department of Pesticide Regulation, Sacramento, CA. Til, H. P., and Beems, R. B. (1979). Sub-acute (4-Week) Oral Toxicity Study with Merpan (Captan) in Rats. Report R-6241, Central Institute for Nutrition and Food Research TNO (MRID 00054467). Til, H., Kuper, C. F., and Folke, H. E. (1983). Life-span Oral Carcinogenicity Study of Merpan in Rats. Report B80–0153, Netherlands Organization for Applied Scientific Research TNO, Zeist, Netherlands (MRID 00161230). Tinston, D. J. (1991). Captan: Teratogenicity Study in the Rabbit. Report CTL/P/3039, ICI Central Toxicology Laboratory, Alderley Park, Macclesfield, UK (MRID 41826901). Tinston, D. J. (1995). Captan: Investigation of Duodenal Hyperplasia in Mice. Report CTL/P/4532, Central Toxicology Laboratory, Alderley Park, Macclesfield, England (MRID 43875602). Tinston, D. J. (1996). Captan: A Time Course Study of Induced Changes in the Small Intestine and Stomach of the Male CD-1 Mouse. Report CTL/P/4893, Central Toxicology Laboratory, Alderley Park, Macclesfield, England (MRID 43982201). Trivedi, S. (1990a). Captan: Excretion and Tissue Retention of a Single Oral Dose (10 mg/kg) in the Rat. Report CTL/P/2820, ICI Central Toxicology Laboratory, Alderley Park, Macclesfield, England (MRID 41505401). Trivedi, S. (1990b). Captan: Excretion and Tissue Retention of a Single Oral Dose (500 mg/kg) in the Rat. Report CTL/P/2862, ICI Central Toxicology Laboratory, Alderley Park, Macclesfield, England (MRID 41505402).

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Trochimowicz, H. J., Kennedy, G. L., and Krivanek, N. D. Jr (2001). Aliphatic and aromatic nitrogen compounds. In “Patty’s Toxicology” (B. Bingham, C. Cohrssen, and C. Powell, eds.) Vol. 4, pp. 1163– 1171. Wiley, New York. U.S. Congress (1996). Food Quality Protection Act of 1996. Pub. Law No. 104–170, 110 Stat. 1989. August 3, 1996. U.S. EPA (1974). Acute Toxicity Screen on PVI Materials and Intermediates. OTS Public Files, Record number 35097, Fiche number: 0000346–0, Document number FYI-OTS-1084–0346, Younger Labs (TSCATS Accession Number 16707). U.S. EPA (1975). Initial Scientific and Minieconomic Review of Captan. Report EPA/540/1–75–012, Office of Pesticide Programs, Criteria and Evaluation Division, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1984a). “Guidance for the Reregistration of Pesticide Products Containing Captafol as the Active Ingredient,” Registration Standard, U.S. Environmental Protection Agency, Office of Pesticide Programs, Washington, DC. U.S. EPA (1984b). Weight of Evidence Classification of Captan as B2. Report 46, Fed. Reg. 46294, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1985a). Captafol: Notice of Special Review. Report 49, Fed. Reg. 1103, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1985b). Captan Position Document 2/3. Report 50, Fed. Reg. 25885, Office of Pesticides and Toxic Substances, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1986a). Classification of Folpet as B2. Report 51, Fed. Reg. 33992, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1986b). “Guidance for the Reregistration of Pesticide Products Containing Captan as the Active Ingredient (EPA Case Number 0120),” U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1986c). Guidelines for Carcinogen Risk Assessment. Report 51, Fed. Reg. 33992–34003, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1987). “Guidance for the Registration of Pesticide Products Containing Folpet as the Active Ingredient (Case Number 0630),” U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1989). Captan; Intent to Cancel Registrations; Conclusion of Special Review (PD4). Report 54, Fed. Reg. 8116–8150, Office of Prevention, Pesticides and Toxic Substances, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1995a). “The HED Chapter of the Re-registration Eligibility Decision Document (RED) for Captan,” Health Effects Division, Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1995b). “The HED Chapter of the Re-registration Eligibility Decision Document (RED) for Folpet,” Health Effects Division, Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1996). Proposed Guidelines for Carcinogen Risk Assessment. Report 61, Fed. Reg. 17960–18011, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1999a). Captan: Reregistration Eligibility Decision (RED). Report 738-R-99–015, Prevention, Pesticides and Toxic Substances (7508C), U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1999b). Folpet: Reregistration Eligibility Decision (RED). Report 738-R-99–011, Prevention, Pesticides and Toxic Substances (7508C), U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1999c). Guidance for Identifying Pesticide Chemicals that Have a Common Mechanism of Toxicity, for Use in Assessing

Chapter | 90  Captan and Folpet

the Cumulative Toxic Effects of Pesticides. Report 6055, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (2004). Captan: Cancer Reclassification; Amendment of Reregistration Eligibility Decision; Notice of Availability. 69 Fed. Reg. 68357-68360 November 24, 2004. U.S. EPA (2007). Pesticide Registration Improvement Renewal Act (PRIA 2) of 2007. U.S. Environmental Protection Agency. Available at: http://www.epa.gov/pesticides/fees/; for fee schedule see: Federal Register: August 5, 2008, 73(151), 45438-45450. U.S. EPA (2008). U.S. EPA Endocrine Disruptor Screening Program Update for the PPDC. Available at: http://www.epa.gov/pesticides/ ppdc/2008/may2008/session4.pdf. Valencia, R. (1981). Mutagenesis Screening of Pesticides Using Drosophila. Report 600/1–81–017, U.S. Environmental Protection Agency, Washington, DC. van Welie, R. T. H., van Duyn, P., Lamme, E. K., Jäger, P., van Baar, B. L. M., and Vermeulen, N. P. E. (1991). Determination of tetrahydrophthalimide and 2-thiothiazolidine-4-carboxylic acid, urinary metabolites of the fungicide captan, in rats and humans. Int. Arch. Occup. Environ. Health 63(3), 181–186. Verma, G., Sharma, N. L., Shanker, V., Mahajan, V. K., and Tegta, G. R. (2007). Pesticide contact dermatitis in fruit and vegetable farmers of Himachal Pradesh (India). Contact Dermatitis 57(5), 316–320. Vogel, E., and Chandler, J. L. R. (1974). Mutagenicity testing of cyclamate and some pesticides in Drosophila melanogaster. Experientia 30(6), 621–623. Vondruska, J. F. (1969). Teratologic Investigation of Captan in Macaca mulatta (Rhesus monkey) and Macac arctoides (Stumptailed

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macaque). Report M5519, Industrial Bio-Test Laboratories, Inc. (MRID 00043398). Vos, J. G., and Krajnc, E. I. (1983). Immunotoxicity of pesticides. Dev. Sci. Practice Toxicol. 11, 229–240. Waner, T. (1988). Folpan: Chronic Oral Study in Beagle Dogs for 52 Weeks. Report MAK.062/FOL, Life Science Research Israel, Ltd., Ness Ziona, Israel. Ward, J. M., Goodman, D. G., Squire, R., Chu, A., and Linhart, M. S. (1979). Neoplastic and nonneoplastic lesions in aging (C57BL/6N  C3H/HeN)F1 (B6C3F1) mice. J. Natl. Cancer Inst. 63, 849–854. Waterson, L. (1995). Folpet: Investigation of the Effect on the Duodenum of MaleMice after Dietary Administration for 28 Days with Recovery. Report MBS 45/943003, Huntingdon Research Centre Ltd. (MRID 44286303). Williams, G. M. (1992). DNA reactive and epigenetic carcinogens. Exp. Toxicol. Pathol. 44, 457–464. Wilson, A., and Wright, A. (1990). A Study of Dermal Penetration of Carbon 14-Folpet in the Rat. Report MAG/1/PH, Toxicol Laboratories, Inc. (MRID 42122018). Wong, Z. A., Bradfield, L. G., and Akins, B. J. (1981). Lifetime Oncogenic Feeding Study of Captan Technical (SX-944) in CD-1 Mice (ICR Derived). Report SOCAL 1150, Chevron Environmental Health Center, Richmond, CA (MRID 00068076). Wong, Z. A., Eisenlord, G. H., and MacGregor, J. (1982). Lifetime Oncogenic Feeding Study of Phaltan Technical (SX-946) in CD-1 (ICR Derived) Mice. Report SOCAL 1331, Chevron Environmental Health Center, Richmond, CA (MRID 125718).

Chapter 91

Mammalian Toxicokinetics and Toxicity of Chlorothalonil P. P. Parsons Syngenta

91.1  Identity and uses of chlorothalonil Chlorothalonil is a halogenated benzonitrile fungicide with broad spectrum activity against vegetable, ornamental, orchard, and turf diseases. It was first registered for use as an agrochemical in the United States in 1966. Chlorothalonil is available in a wide variety of formulations including suspension concentrates, wettable powders, and water dispersible granules. Its mode of fungicidal action is to bind to sulfhydryl groups of amino acids, proteins, and peptides and, in doing so, it ties up free glutathione in fungal cells, thereby blocking glycolytic and respiratory enzyme pathways. This action prevents the ability of fungal cells to infect plants and results in death of the fungus. Chlorothalonil’s multisite mode of action has meant that no significant problem with fungal resistance has been encountered. In addition to its use as an agricultural fungicide, chlorothalonil also has wider biocidal applications, for example in paints and lubricating fluids.

Chemical class Empirical formula Synonyms and trade names

halogenated benzonitrile CCl4N2 Bravo®, Daconil®, Tuffcide®, Acticide®

91.1.1.3  Physical Properties Physical appearance Solubility (at 25°C)

Vapor pressure Molecular weight Melting point log Po/w

grey/white crystalline solid, odorless in pure form practically insoluble in water (0.6–0.8 mg/l) xylene—80 g/l acetone—20 g/l cyclohexane—30 g/l 7.62  105 Pa at 25°C 265.9 250–251°C 2.94 at 25°C

91.2  Mammalian toxicokinetics 91.1.1  Physical and Chemical Properties 91.1.1.1  Structure CN



CI

CI

CI

CN CI

91.1.1.2  Chemical Identity Common name CAS No. EINECS Chemical name

chlorothalonil 1897-45-6 217-588-1 2,4,5,6tetrachloroisophthalonitrile

Hayes’ Handbook of Pesticide Toxicology Copyright © 2001 Elsevier Inc. All rights reserved

An overview of the available metabolism and pharmacokinetic data for chlorothalonil has been published by Wilkinson and Killeen (1996).

91.2.1  Oral Administration In rats, given a single, oral low dose (1.5 mg/kg) of chlorothalonil, around 20–22% of the absorbed dose is excreted in bile and around 10% in urine (Marciniszyn et al., 1985a, b, 1986a). At higher doses (200 mg/kg) a considerably lower proportion (8%) of the absorbed dose is excreted in bile, indicating that this is a saturable process. These data indicate that overall absorption from the G.I. tract is in the order of 30–32% of the administered dose. The majority of radiolabel is excreted in feces with at least 80% of administered dose 1951

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excreted by this route within 96 h. Approximately 90% of the administered dose was excreted within 24–48 h although excretion was less rapid at doses of 50 mg/kg and above. Highest tissue concentrations were observed in the kidney, approximately 0.1% of the dose. A similar metabolic profile was seen on repeated dosing and there was no evidence for bioaccumulation (Savides et al., 1986a, b). Thiol-derived metabolites were identified in urine. Following administration of similar doses of chlorothalonil to germ-free rats, only 3% of the dose appeared in urine with lower proportions excreted as thiol-derived metabolites, indicating that gut microflora may play a role in the disposition and metabolism of chlorothalonil in the rat. Bile cannulation studies have confirmed that chlorothalonil undergoes enterohepatic circulation in the rat (Marciniszyn et al., 1986b). In dogs, approximately 6% of an oral dose of 50 mg/ kg was excreted within 48 h (1% in urine and 5% in bile). As with the rat, absorption and subsequent excretion were rapid with around 89% of an administered dose recovered within 48 h. The extent of urinary thiol-derived metabolite excretion in dogs was lower than that seen in rats. At necropsy at 48 h approximately 0.1% of the administered dose was present in the liver and kidneys with 0.01% in other tissues (Savides et al., 1995). Limited data for the monkey show that, following a single oral dose of 50 mg/kg, 1.8–4.1% of the dose appeared in urine with very low levels of thiol-derived metabolites appearing in urine. Fecal excretion predominated with around 92% of the dose eliminated via this route over 96 h. Absorption and excretion were rapid and there was no evidence of bioaccumulation (Savides et al., 1990).

There are limited data concerning the disposition and excretion of chlorothalonil in the mouse with no metabolism data in this species. Low levels of radioactivity were found in the tissues and urinary excretion indicated that at least 10% of the dose was absorbed with the majority (70– 80%) of the dose excreted in faeces (Ribovich et al., 1982). Comparison of the differences in urinary metabolite excretion profile between species suggests that the ability to excrete thiol-derived metabolites may be correlated with the observed species differences in susceptibility to renal toxicity. Mechanistic studies have been conducted to determine if a relationship exists between the ability to excrete urinary thiol-derived metabolites of chlorothalonil and the potential to induce renal toxicity. Inhibition of   -glutamyltranspeptidase using Acivicin (Savides et al., 1985) and renal organic anion transport using probenecid (Marciniszyn et al., 1986b) decreased urinary thiol-derived metabolite excretion in rats. Administration of the monoglutathione conjugate to rats was shown to produce a qualitatively similar pattern of metabolite excretion to that seen following administration of chlorothalonil itself (Mead et al., 1987a, b). These studies indicate that excretion of thiol-derived metabolites of chlorothalonil requires glutathione conjugation and then subsequent enzymatic processing of glutathione-derived conjugates that are selectively accumulated within the kidney. By analogy with other chemicals that undergo extensive glutathione conjugation, it is reasonable to presume that metabolism proceeds via cysteine conjugates and N-acetyl cysteine conjugates (“mercapturates”) as outlined in Fig. 91.1.

Figure 91.1  Diagram illustrating proposed metabolism of chlorothalonil following oral administration to rats. Broken lines indicate multistage events involving several enzymatic steps and transport processes.

CN CI

CI

CI

CN CI

Di- & tri-glutathione conjugates

Thil-derived metabolites in plasma

Biliary excretion Di- and tri-glutathione conjugates in bile

Further processing

Chlorothalonil & metabolites excreted in feces

Enterohepatic re-circulation

Thiol-derived metabolites in kidney

Thiol-derived metabolites excreted in urine

Chapter | 91  Mammalian Toxicokinetics and Toxicity of Chlorothalonil

In conclusion, data from a variety of species demonstrate that, following oral administration, chlorothalonil is rapidly absorbed with fecal excretion predominating. The toxicokinetic profile is similar on repeated dosing with no evidence for bioaccumulation. Glutathione conjugation plays a central role in the metabolism of chlorothalonil and subsequent complex metabolic processing of these conjugates results in selective renal uptake and urinary excretion of thiol-derived metabolites. Knowledge from the metabolism of other chemicals that undergo extensive glutathione conjugation implicates a role for mercapturic acidmediated metabolism for chlorothalonil.

91.2.2  Dermal Administration Studies have been conducted to determine the nature of metabolites appearing after dermal administration of chlorothalonil to the rat and monkey. Separate in vitro and in vivo studies have been conducted to determine the extent to which chlorothalonil undergoes percutaneous absorption.

91.2.2.1  Urinary Metabolite Profile Following Dermal Exposure Limited studies have been conducted in the rat and the monkey to investigate the profile of urinary metabolite excretion after dermal administration of chlorothalonil. In rats, a maximum of 3.1% of the applied dose was excreted in urine of which around 0.1–1% constituted thiol-derived metabolites (Savides et al., 1989). Fecal excretion of radiolabel was similarly low. A high proportion (20–40%) of administered dose was retained in skin at the application site. The low level of thiol-derived metabolites excreted in urine following dermal administration of chlorothalonil may explain the absence of renal toxicity seen in the subchronic toxicity studies using this route. In monkeys, only 1.2% of the dose was excreted in urine and similar amounts in feces. Around 2–4% of the dose was retained in skin at the application site after a 48 h exposure and no thiol metabolites could be detected in urine (Magee et al., 1990).

91.2.2.2  Percutaneous Absorption A number of percutaneous absorption studies have been conducted with chlorothalonil using both in vitro and in vivo approaches. These studies used either acetone or formulation blank as the vehicle. In Vitro Studies Two in vitro studies have been conducted using human abdominal epidermis. In one study, the chlorothalonil was used either neat (no vehicle) or as a solution in acetone (Ward, 1989a). The mean absorption rate observed in this study over a 48 to 55 h exposure interval was 0.034  0.020 g/cm2/h, equating to 0.085% of the applied

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dose. No radiolabel was detected in the receptor chamber fluid until 48 h after application using either neat or diluted material. In a separate study (Ward, 1989b), chlorothalonil was applied as a suspension in a commercial formulation base either as a concentrate or as a spray-strength dilution under both occluded and nonoccluded conditions. The mean absorption rate per hour was higher with the occluded application than with the nonoccluded application (0.18 vs 0.005 g/cm2/h). Approximately 0.094% of the applied dose was absorbed over a 10 h period with the nonoccluded application. These data indicate that the absorption of chlorothalonil through the human epidermis is considerably lower than 1% of the applied dose. In Vivo Studies A study was conducted in the rat which investigated the percutaneous absorption of chlorothalonil using acetone and a blank commercial formulation as dosing vehicles (Andre et al., 1991a). Absorption values ranging from 6 to 26% were obtained taking into account the radiolabelled material bound to skin at the application site. A more realistic indication of systemic absorption is obtained when skin bound material is discounted. Using this approach, percutaneous absorption at 10 h postapplication in formulation blank was 2.2, 3.6, and 0.4% of the applied using applications of 0.1, 0.5, and 5.0 mg/kg, respectively. These values are consistent with the absorption profile seen in vitro. Thus, the available data indicate that chlorothalonil is a poor skin penetrant. This view is supported by the estimation of percutaneous absorption by comparison of the ability of chlorothalonil to induce toxicity within the key target organ (kidney) following dermal and oral administration to rats. Such a comparison is based on the findings of the 21 day dermal toxicity study (see Section 91.5.2) with the interim findings after a similar dosing period in a 90 day subchronic oral toxicity study (see Section 91.5.1). In these respective studies, the NOAELs for the induction of renal tubular hyperplasia were 600 mg/kg/d (the highest dose) and 3 mg/kg/d. Toxicokinetic studies in rats indicate that approximately 32% of an oral dose of chlorothalonil is absorbed from the G.I. tract. Therefore, the NOAEL for renal hyperplasia after 6 weeks in the 90 day oral rat study of 3 mg/kg/d equates to a systemic dose of 0.96 mg/kg/d (i.e., 0.32  3). Since the application of 600 mg/kg/d of chlorothalonil to rat skin did not result in any kidney toxicity, it can be deduced that that approximately 0.16% of the applied dermal dose was absorbed systemically (i.e., 0.96/600  100%).

91.3  Acute toxicity 91.3.1  Oral Chlorothalonil is not acutely toxic by the oral route having a maximum lethal dose (MLD)  5000 mg/kg in the rat

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with no deaths at 5000 mg/kg in carbosa methyl cellulose (Moore, 2000). The only clinical signs of toxicity were congenital staining, soft feces, and/or occurred on the day of dosing and the day following dosing.

91.3.2  Dermal The dermal MLD was 10,000 mg/kg in rabbits with no deaths observed at this dose. In this study, chlorothalonil was applied to abraded skin for 24 h (Shults et al., 1981b). Slight edema and yellow discoloration were observed at the application site and eye irritation was also seen. At necropsy, pale areas were observed in the liver. Chlorothalonil is not systemically toxic by the dermal route.

91.3.3  Intraperitoneal The intraperitoneal MLD has been estimated to be 3.2 mg/ kg in the rat. Clinical signs of toxicity were not reported. Pathological findings were generally consistent with injection of an irritant substance into the peritoneal cavity with chronic fibrinous peritonitis, enlarged and congested mesenteric lymph nodes, intestinal edema, and red foci on the kidneys and lungs (Wazeter and Lucas, 1971).

91.3.4  Inhalation Acute inhalation exposure (whole body) of rats to an atmosphere containing chlorothalonil dust resulted in a 4 h MLC of 0.1 mg/l (Shults et al., 1993). Mortality was dose-related with deaths at 0.08, 0.14, and 0.21 mg/l (4, 6, and 9/10 animals, respectively) occurring from 10 minutes to 2 days postexposure. Clinical signs of toxicity included gasping, eye closure, and exaggerated breathing observed during exposure. Recovery was evident from day 4 onward and the majority of animals were normal by day 9. Congestion of the lungs and white frothy fluid in the trachea were observed in those animals that died with no abnormal pathology in surviving animals. The MMAD was 2.5–3.6 m with 25% of particles 2 m in diameter. In a separate study (Shults et al., 1981c), rats were exposed (whole body) to a dust atmosphere containing chlorothalonil at concentrations of 0.07–0.22 mg/1 for 4 h. The MMAD was 1.35–5.5 m with 90% of particles of 10 m in diameter. The MLC was estimated to be 0.09 mg/l with a dose-related incidence of mortality (1/20 at 0.7 mg/l vs 19/20 at 0.22 mg/l). Clinical signs of toxicity included rales and a bloody nasal discharge. Pulmonary congestion was observed in animals that died during the study. Hepatic necrosis and deposition of eosinophillic material were observed at the top dose. A high incidence of respiratory mycoplasmosis was seen in all animals in this study. Although the study is compromised by the presence

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of concomitant infection, the 4 h MLC was in agreement with that observed in the study above. It is concluded that chlorothalonil is toxic by inhalation causing death by asphyxiation secondary to pulmonary edema. In standard regulatory studies, the clinical signs of toxicity and pathological findings are consistent with exposure to a substance that causes pulmonary irritation to the lungs and respiratory tract.

91.3.5  Skin Irritation Technical chlorothalonil is not a skin irritant in rabbits (Shults et al., 1981d). In this study, a prolonged exposure period of 24 h was used and effects were studied using both abraded and intact skin. Although the study conditions were designed to maximize the potential to induce skin irritation, the only effects observed were isolated signs of mild irritation. In contrast, prolonged and/or repeated dermal exposure to chlorothalonil has been shown to produce significant signs of skin irritation in acute and subchronic dermal toxicity studies in the rat and rabbit (see Section 91.5.2). It is concluded that, while chlorothalonil is not a skin irritant in standard studies for assessment of this endpoint, it does display potential to cause skin irritation in other dermal toxicity studies involving prolonged or repeated dermal application. Chlorothalonil has been shown to cause dermal reactions in humans who have been occupationally exposed to chlorothalonil, although it has not always been apparent if this constitutes and irritation or sensitization response (see Section 91.11.1).

91.3.6  Eye Irritation In a standard rabbit eye irritation study, chlorothalonil caused irreversible ocular lesions in rabbits with corneal opacity persisting for up to 14 days postinstillation. Effects also persist in the iris and conjunctiva (Wilson, 1977a). A further four studies have been conducted with technical chlorothalonil and each of these studies demonstrated irreversible corneal opacity (Francis et al., 1973; O’Meara and Laveglia, 1995; Wilson, 1977b, c). The severity and persistence of these eye lesions indicates that chlorothalonil has potential to cause serious damage to eyes. The effect of washing the eyes postcompound instillation has not been investigated using the technical material itself, although studies with high strength formulations have shown that postinstillation washing ameliorates these effects. This observation is relevant to the recommended treatment of individuals following accidental ocular exposure. Experience from accidental human exposure indicates that chlorothalonil causes ocular pain and is also irritating to the human eye (see Section 91.11.2). However, the severe

Chapter | 91  Mammalian Toxicokinetics and Toxicity of Chlorothalonil

and irreversible eye lesions seen in the rabbit have not been documented in humans.

91.3.7  Summary of Acute Toxicity Chlorothalonil has very low acute toxicity by the oral and dermal routes although it is very toxic by inhalation. Many of the effects seen following acute exposure are consistent with irritation at the initial site of contact. However, chlorothalonil was not irritating to skin when tested in a standard skin irritation study although dermal irritation has been observed in acute and subchronic toxicity studies in the rat and rabbit, indicating the potential for chlorothalonil to cause skin irritation following repeated or prolonged exposure. Chlorothalonil causes irreversible and severe ocular lesions in rabbits. The acute toxicity of chlorothalonil is summarized in Table 91.1.

91.4  Sensitization

1955

Table 91.1  Acute Toxicity of Chlorothalonil Study

Species

Result

Oral toxicity

Rat

MLD  10,000 mg/kg

Dermal toxicity

Rabbit

MLD  10,000 mg/kg

Inhalation toxicity

Rat Rat Rat

4 h MLC 0.1 mg/l 4 h MLC 0.09 mg/1 1 h MLC 0.52 mg/1

Rat Rat Mouse

MLD 3.2 mg/kg MLD 8 mg/kg MLD 12 mg/kg

Skin irritation

Rabbit

aNot a skin irritant

Eye irritation

Rabbit Rabbit Rabbit Rabbit Rabbit

Irreversible corneal opacity Irreversible corneal opacity Irreversible corneal opacity Irreversible corneal opacity Irreversible corneal opacity

Intraperitoneal toxicity

a

Potential skin irritant with prolonged or repeated dermal exposure.

91.4.1  Skin Sensitization The potential for chlorothalonil to induce skin sensitization has been investigated in guinea pigs using a number of different study designs (see Table 91.2). Although these studies provide an inconsistent profile with regard to skin sensitization potential, the data are supportive of the view that chlorothalonil is a skin sensitizer of relatively low potency. It appears that a concentration of at least 40% w/v technical chlorothalonil is required for the induction of sensitization in guinea pigs. It is concluded that chlorothalonil is a weak skin sensitizer in guinea pigs. In addition to these data in animals, information is available concerning the potential of chlorothalonil to cause skin sensitization in humans following dermal exposure in the occupational setting (see Section 91.11.1).

91.4.2  Respiratory Sensitization There are no animal data concerning the potential for chlorothalonil to induce respiratory sensitization.

91.5  Subchronic toxicity 91.5.1  Oral In rats, dietary administration of chlorothalonil for 28 days caused clinical signs of toxicity, decreased body weight, and decreased hematological parameters at doses of 375 mg/ kg/d. One death occurred at the top dose (1500 mg/kg/d). Absolute kidney weight was increased at 175 mg/kg/d. No effects were observed at 80 mg/kg/d but this cannot reliably be considered as a NOEL as histopathological examination was not conducted in this study and longer

term studies suggest that hyperplasia of the forestomach and proximal tubular epithelium can occur at this dose level (Wilson et al., 1982b). The principal lesions observed following dietary administration of chlorothalonil to rats and mice for up to 90 days were hyperplasia and hyperkeratosis of the forestomach and hyperplasia of the proximal tubular epithelium of the kidney (Shults et al., 1983, 1985; Wilson et al., 1983a, b, 1984, 1985a, b). No treatment-related mortality was observed in these studies at doses up to 1500 mg/kg/d and there were only limited clinical signs of toxicity. Renal hyperplasia only occurred at a low incidence at the top dose tested in male mice with no effect in females. Other effects included decreased plasma ALT activity and increased kidney weight. The overall NOELs in these studies were 1.5 mg/kg/d for rats and 2.8 mg/kg/d for mice with respective LOELs of 3.0 and 9.2 mg/kg/d. The NOEL for hyperplasia of the forestomach was 3 mg/kg/d in both species with LOELs of 10 mg/kg/d in the rat and 9 mg/kg/d in the mouse. The forestomach lesions were fully reversible after a 13 week recovery period in rats but were not investigated in mice. The NOAEL for proximal tubular hyperplasia after 6 weeks dosing was 3 mg/kg/d in rats with a clear increase in the incidence of hyperplasia at 40 mg/kg/d in animals necropsied at 13 weeks. Proximal tubular hyperplasia was evident in 2/5 animals at 6 weeks at 10 mg/kg/d. In male mice, the NOEL for renal hyperplasia was 48 mg/kg/d with a LOEL of 130 mg/kg/d. Further investigative studies have been conducted in the rat. Vacuolar degeneration has been shown to be present

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Table 91.2  Summary of Skin Sensitization Studies Conducted in Guinea Pigs Test method (date)

Induction concentration

Challenge concentration

Conclusion

Reference

9 induction Buehler (1974)

10% in water

10% in water

Not a sensitizer

CTL/C/3873 Shults and Wilson, (1985)

9 induction Buehler (1985)

100%

10% in saline

a

9 induction Buehler (1985)

100%

100%

a

CTL/C/3573

Maximization (1986)

i.d.—0.5% in acetone topical—1% in acetone

0.001% in acetone

a

Tucker (1986)

Maximization (1986)

i.d.—5% in propylene glycol topical—1% in acetone

0.0125% in acetone

Sensitizer

Tucker (1986)

20 induction open cutaneous (1986)

0.01, 0.1, or 1%

0.0025, 0.0075 or 0.025%

Sensitizer

Tucker (1986)

10-induction Buehler (1982)

100%

100%

Not a sensitizer

Wilson et al. (1982a)

Maximization (1988)

Acetone vehicle

Acetone vehicle

Equivocal

Wilson et al. (1988a)

Buehler

Acetone vehicle

Acetone vehicle

Sensitizer

Wilson et al. (1988b)

Maximization

Acetone vehicle

Acetone vehicle

Sensitizer

Wilson et al. (1988c)

Sensitizer Not a sensitizer Not a sensitizer

a

Result confirmed by rechallenge.

in the proximal tubular epithelium after only two daily doses of chlorothalonil at 175 mg/kg/d (Ford et al., 1988). In 28 day (Hironaka et al., 1996) and 90 day (Mizens et al., 1996) studies, chlorothalonil has also been shown to increase cell proliferation in the forestomach (BrdU) and proximal tubule (PCNA) of rats. Significant increases in labelling indices were evident from day 7 of dosing through to days 28 or 90 respectively at doses of 15 mg/ kg/d. The NOEL for increased labelling indices in both tissues was 1.5 mg/kg/d. Similar morphological lesions were seen in the renal proximal tubular epithelium following gavage administration of equivalent doses of the monoglutathione conjugate of chlorothalonil (150 mg/kg/d) or parent compound (75 mg/kg/d), although in these animals no hyperplastic changes were seen in the forestomach (Mead, 1987b). This finding implicates glutathione conjugation in the metabolism of chlorothalonil-induced renal toxicity and demonstrates that it is the parent compound that causes toxicity in the rodent forestomach. In dogs, administration of chlorothalonil for 90 days caused a decrease in body weight gain at 150 and 500 mg/ kg/d with one death seen at the top dose (Fillmore et al., 1993). There was some indication of a decrease in mean body weight throughout the study in both sexes at 15 mg/ kg/d, but this was not considered to be toxicologically significant. Some changes in clinical chemistry were seen at these dose levels with plasma ALT decreased at all dose levels. The NOAEL was 15 mg/kg/d as effects on ALT

were not considered adverse. No lesions were observed in the stomach or the kidney, indicating that the dog has a different toxicity profile to rodents.

91.5.2  Dermal Subchronic (21 day) dermal toxicity studies have been conducted in rats and rabbits. In rabbits, no systemic effects were observed that were associated with chlorothalonil administration at doses up to 50 mg/kg/d (Shults et al., 1986). However, histopathological examination revealed evidence of parasitic infection in all animals which compromises the value of this study. The NOEL for local effects in the skin was 0.1 mg/kg/d based on the observation of erythema and skin thickening at 2.5 mg/kg/d. In rats, decreases in body weight gain were observed over the 21 day study period along with simultaneous decreases in food consumption (Mizens et al., 1986a). The decrease in weight gain was more prominent early in the study with increasing weight gain observed toward the end of the study. Plasma ALT activity was decreased at all dose levels. No histopathological effects were observed in the kidney, indicating that the NOEL for renal effects following dermal application was 600 mg/kg/d. Erythema, hyperkeratosis, and squamous epithelial cell hyperplasia where observed at the site of application at all dose levels, indicating the potential for chlorothalonil to cause dermal irritation with repeated exposure.

Chapter | 91  Mammalian Toxicokinetics and Toxicity of Chlorothalonil

1957

Table 91.3  Summary of Subchronic Toxicity Data Species

Route

Duration (days)

NOEL (mg/kg/d)

LOEL (mg/kg/d)

Rat

Oral diet

90

None

40 (hyperplasia of forestomach and kidney)

Rat

Oral diet

90

1.5 3 10

3 (↑ kidney weight, ↓ ALT) 10 (hyperplasia of forestomach) 40 (hyperplasia of kidney)

Mouse

Oral diet

90

3 3 48

9 (↑ kidney weight, ↓ ALT) 9 (hyperplasia of forestomach) 130 (hyperplasia of kidney)

Dog

Oral—capsule

90

15

150 (↓ weight gain)

Rabbit

Dermal

21

0.1 50

2.5 local effects 2.6 no systemic effects

Rat

Dermal

21

None 600

60 local effects No renal histopathology observed at top dose—600

91.5.3  Inhalation No repeat-dose inhalation toxicity studies have been conducted with technical chlorothalonil. However, the symptoms and pathological findings seen after acute inhalation exposure are confined to the respiratory tract with no evidence of systemic toxicity. It is therefore anticipated that toxicity in a repeat-dose inhalation study would be expected to manifest as local irritation with the NOAEL being driven by site-of-contact toxicity within the lungs and respiratory tract as opposed to systemic toxicity. The subchronic toxicity of chlorothalonil is summarized in Table 91.3.

91.6  Chronic Toxicity The chronic oral toxicity of chlorothalonil has been determined in the rat and mouse with two long-term studies available in each species. The initial studies on chlorothalonil (Wilson et al., 1983c, 1985c) were conducted at relatively high dose levels and failed to demonstrate NOELs, so the studies were repeated (Wilson et al., 1987, 1989) using lower doses. The findings in the chronic rat and mouse studies were consistent with those seen in the subchronic studies with hyperplasia of the forestomach and proximal tubular epithelium being the most prominent effects. The hyperplastic changes in the proximal tubular epithelium were associated with an increase in absolute kidney weight. The NOELs for hyperplasia and hyperkeratosis of the forestomach were 1.8 and 1.6 mg/kg/d for rats and mice respectively and the NOELs for hyperplastic changes in the proximal tubular epithelium were 1.8 mg/ kg/d for rats and 5.4 mg/kg/d for mice. In addition to the

hyperplastic lesions in the kidney and forestomach, some minor changes were seen in clinical chemistry, hematology, and urinalysis. There was no evidence of systemic toxicity in organs other than the kidney and forestomach. Although changes occurred in some organ weight: body weight ratios, these were attributable to decreases in body weight. In beagle dogs, 1 year administration of chlorothalonil caused a significant decrease in body weight gain at 500 mg/kg/d with increases in absolute liver and kidney weights at 150 mg/kg/d (Mizens and Laveglia, 1994). No histopathological findings were seen in association with the organ weight changes. Changes in clinical chemistry were seen but only at one time point (27 weeks) and with no relationship to dose. Increased pigmentation of the kidney was observed at 150 and 500 mg/kg/d. The severity of this effect was similar at both dose levels and minimal pigmentation was seen in control and low dose animals. Due to its presence in control animals and lack of relationship to dose, the renal pigmentation was not considered to be of toxicological significance. Moreover, this observation was clearly distinct from the hyperplastic lesion seen in the renal proximal tubular epithelium of rodents. This indicates that there are species-specific differences in susceptibility to the renal toxicity of chlorothalonil. Although dogs do not possess an anatomical equivalent of the rodent forestomach, the stomachs from this study were examined to determine if there was any evidence of cell proliferation using PCNA labelling. There was no increase in labelling index in any treated group when compared to concurrent controls, providing reassurance that chlorothalonil is not toxic to the gastric mucosa of dogs. The overall NOAEL for this study was 150 mg/kg/d.

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Table 91.4  Summary of Chronic Toxicity Data Duration of study

Species

NOAEL (mg/kg bwt/day)

LOEL (mg/kg bwt/day)

2 years

Rat

None

40

2 years

Rat

3.8 1.8 1.8

15 (kidney tumors) 3.8 (kidney hyperplasia) 3.8 (stomach mucosal tumors) 3.8 (stomach squamous cell tumors)

1.8 2 years

Mouse

None

125

2 years

Mouse

5.4 23

23 (kidney hyperplasia) 100 (stomach mucosal tumors) 5.4 (stomach squamous cell tumors)

1.6 1 year oral

Dog

150

500 (decreased body weight gain)

The chronic toxicity findings are summarized in Table 91.4.

91.7  Genotoxicity 91.7.1  In Vitro Genotoxicity Studies An extensive range of tests have been conducted to assess the genotoxic potential of technical chlorothalonil. In vitro studies were mostly negative including Ames tests using renal and hepatic metabolic activation systems. Up to 17 metabolites of chlorothalonil have also been tested and shown to give gave negative results using rat kidney S9. Two nonstandard bacterial DNA repair assays have been performed, one of which was positive and the other negative. A significant increase in the incidence of chromosomal aberrations has been observed in Chinese hamster ovary cells although this was only evident in the absence of auxiliary metabolic activation (Mizens et al., 1986b). In the absence of S9, increases in structural aberrations over control values were observed only at the top two concentrations which approached cytotoxic levels. These data indicate that chlorothalonil has clastogenic activity in vitro in the absence of S9. The effect was not observed in the presence of S9, even at dose levels some 20 times higher. However, given the lack of genotoxicity observed in other in vitro test systems (e.g., the Ames test) and the known reactivity of chlorothalonil, a rationale for such a profile is possible. Chlorothalonil can be viewed as a reactive molecule insofar as it is known to be reactive towards thiol (-SH) groups. It can be considered as a soft electrophile with a

preference for sulphur nucleophiles rather then nitrogen/ oxygen nucleophiles. Such chemicals tend to show reactivity toward protein (contains critical S electrophiles) rather than toward DNA (contains critical O and N nucleophiles). The profile of activity in genotoxicity assays is that such material displays negative findings in the Ames test but apparent positive findings in the in vitro cytogenetics assay, usually in the absence of exogenous metabolic activation. The activity of chlorothalonil in the IVC assay is likely to be through reactivity with protein (not DNA), and with the protein dependency of the chromosomal structure allowing visualization as a structural aberration. Such an activity would not be expected to produce genotoxicity in vivo, as reaction with inter- and intracellular thiols would dissipate the activity. This is supported by the observation that the in vitro activity of chlorothalonil in the IVC assay is removed by the addition of S9, and also by the fact that in vivo metabolism/distribution studies have confirmed that chlorothalonil reacts very rapidly with thiols. The clastogenic response observed in vitro in the IVC assay is therefore considered to be of no real significance regarding possible genotoxicity in vivo especially when considered in light of the findings of the in vivo cytogenetic studies with chlorothalonil.

91.7.2  In Vivo Genotoxicity Studies Several in vivo bone marrow cytogenetic studies have been conducted with chlorothalonil using single or repeated dosing schedules in three different species (rat, mouse, and Chinese hamster). Most of these studies used high doses of chlorothalonil (up to 5000 mg/kg) which resulted in some mortality. All of these in vivo cytogenetic studies were negative indicating that the clastogenicity seen with chlorothalonil in vitro is not manifest in vivo. Further reassurance for lack of genotoxic activity in vivo comes from a study which demonstrated that intraperitoneal administration of radiolabelled chlorothalonil to the rat did not result in any labelled material binding covalently to rat kidney DNA. It is therefore concluded that chlorothalonil is not genotoxic in vivo. The results of the genotoxicity studies conducted with chlorothalonil are summarized in Table 91.5.

91.8  Carcinogenicity Treatment-related increases in the incidence of renal tubular adenoma and carcinoma were observed in rats and male mice (Wilson et al., 1983c, 1985c, 1987, 1989). Squamous cell adenomas and carcinomas were also observed in the forestomach of both species. In dogs, there was no evidence of neoplastic development nor was there any evidence for the occurrence of preneoplastic lesions in the kidney or stomach after administration of chlorothalonil

Chapter | 91  Mammalian Toxicokinetics and Toxicity of Chlorothalonil

1959

Table 91.5  Summary of Genotoxicity Data Test

Study type

Result

Reference

In vitro

Ames test (hepatic activation) Ames test (renal activation) Chromosomal aberration assay    (CHO) Gene mutation assay (Chinese hamster V-79 cells & mouse fibroblast BALB/3T3 cells) DNA repair test (S. typhimurium) DNA repair test (B. subtilus) Cell transformation assay (F1706 P95 and H4536 P97 cells)

Negative ( activation) Negative ( activation) Negative ( activation) Positive ( activation) Negative ( activation)

Banzer and Kouri (1977) Jones et al. (1984) Mizens et al. (1986b)

Positive Negative Negative

Auletta and Kouri (1977) Shirasu (1978) Price and Bailee (1979)

Micronucleus test

Negative (rat) Negative (mouse) Negative (Chinese hamster) Negative (rat) Negative (mouse) Negative (Chinese hamster) Acute study—negative Subchronic study—negative Negative (rat) Negative (Chinese hamster)

Killeen and Siou (1983)

In vivo

Chromosome aberration test

Chromosome aberration test (Chinese hamster) Covalent binding to DNA Chromosome aberration test

for up to 1 year. The NOEL for mucosal cell tumors of the forestomach was 1.8 mg/kg/d in rats and 23 mg/kg/d in mice. In rats, the NOEL for tumors of the proximal tubular epithelium was 3.8 mg/kg/d in males and 15 mg/kg/d in females. In the first mouse carcinogenicity study, renal tubular adenomas and carcinomas were observed at all dose levels in males including the lowest dose of 125 mg/kg/d. In a subsequent study using the same strain of male mice, there were no treatment-related increases in the incidence of renal tumors up to a dose level of 100 mg/kg/d. Table 91.6 summarizes the key NOAELs observed in these studies.

Killeen (1983)

Siou et al. (1985) Savides et al. (1987) Proudlock (1995)

Table 91.6  Summary of Carcinogenicity Findings Duration of study

Species NOAEL (mg/kg bwt/day)

LOEL (mg/kg bwt/day)

2 years

Rat

None

40

2 years

Rat

3.8 1.8 1.8

15 (kidney tumors) 3.8 (kidney hyperplasia) 3.8 (stomach mucosal tumors)

2 years

Mouse

None

125

2 years

Mouse

5.4 23

23 (kidney hyperplasia) 100 (stomach mucosal tumors)

150

500 (decreased body weight gain)

91.8.1  Mode of Carcinogenic Action A mechanistic interpretation for the carcinogenicity of chlorothalonil has been published by Wilkinson and Killeen (1996).

Kouri (1977)

1 year oral Dog

91.8.1.1  Forestomach Tumors Repeated administration of chlorothalonil causes hyperplasia in the forestomach of rats and mice. The data are consistent with a temporal sequence of events starting with increased cell proliferation, multifocal ulceration and erosion of the forestomach mucosa, regenerative hyperplasia and hyperkeratosis, ultimately progressing to the formation of gastric tumors within the forestomach. Clear thresholds have been demonstrated for the induction of hyperplasia

and neoplasia in both species. The fact that chlorothalonil is not genotoxic provides reassurance that these tumors occur as a secondary consequence of local irritation within the rodent forestomach. Oral subchronic studies with the monoglutathione conjugate of chlorothalonil, failed to induce any toxicity in the rat forestomach, indicating that it is parent chlorothalonil, and not a metabolite, that is the toxic agent at this site.

1960

In contrast to the findings in the rat, there was no evidence of preneoplastic stomach lesions in dogs orally administered chlorothalonil for up to 1 year at dose up to 500 mg/kg/d, a dose level considerably higher than that which causes hyperkeratosis and hyperplasia in the rodent forestomach (approximately 4 mg/kg/d in the rat). The absence of any evidence of increased cell proliferation in the dog stomach was confirmed by PCNA labelling of tissue obtained at the termination of this study. The absence of stomach lesions in the dog is attributable to the anatomical differences between rodents and dogs in that dogs do not possess a forestomach. Similarly, humans are like the dog in that they do not possess an anatomical equivalent of the rodent forestomach. It is therefore concluded that the rodent forestomach tumors induced by chlorothalonil are not indicative of a carcinogenic risk to humans.

91.8.1.2  Renal Tumors The experimental data show a temporal sequence of events that lead to the formation of renal tumors in rats and male mice. Vacuolar degeneration of the renal proximal tubular epithelium has been shown to occur after two daily doses of chlorothalonil and cytotoxicity and degeneration of tubular epithelial cells can be seen after only 2 days treatment. Continued administration of chlorothalonil leads to the development of a regenerative hyperplasia within the renal proximal tubular epithelium. Continued regenerative hyperplasia ultimately results in progression of the kidney lesion to tubular adenoma and carcinoma. Thus, the data clearly show that initial cytotoxicity and regenerative hyperplasia within the proximal tubular epithelium are essential prerequisites for subsequent tumor development. Clear thresholds have been demonstrated for this nongenotoxic secondary mode of action which is a direct consequence of chronic stimulation of cell proliferation. Doses of chlorothalonil below the threshold for the induction of these preneoplastic lesions would not be expected to be carcinogenic. Studies have been conducted investigating the role of the glutathione conjugation pathway and subsequent formation of urinary thiol-derived metabolites in renal tumor formation. The central role of glutathione in the metabolism and subsequent toxicity of chlorothalonil has been shown by studies with a monoglutathione conjugate of chlorothalonil. Knowledge of the metabolism of other chemicals that undergo extensive glutathione conjugation implicates a role for mercapturic acid-mediated metabolism for chlorothalonil. Maneuvers that inhibit key enzymes in this metabolic process, such as inhibition of the activity of glutamyltranspeptidase or the renal organic anion transporter, decrease the urinary excretion of thiol-derived metabolites in the rat. Administration of a monoglutathione conjugate to rats caused similar lesions in the kidney to parent chlorothalonil

Hayes’ Handbook of Pesticide Toxicology

although no effects were seen in the forestomach. Studies undertaken in vitro using isolated kidney mitochondria have shown that respiration is inhibited in the presence of synthetic mono- and dithiol conjugates derived from chlorothalonil (Andre et al., 1991b; Savides et al., 1988). Furthermore, a correlation appears to exist between the interspecies differences in susceptibility to renal toxicity and the differences in capacity to produce these thiol-derived metabolites as rats excrete more thiol-derived metabolites than dogs. The proposed mode of action for the induction of renal toxicity in rodents is outlined in Fig. 91.2. It is concluded that chlorothalonil is a nongenotoxic kidney carcinogen in rats and mice and the NOAELs observed for both tumors and the precursor lesions indicate that it is appropriate to assume that a threshold exists for carcinogenicity. The species differences in metabolism are reflected in the different toxicity profiles seen in rodents and dogs and suggest that the dog is the most appropriate species for human health risk assessment and that these tumors are highly unlikely to develop in humans exposed to chlorothalonil.

91.9  Reproductive Toxicity 91.9.1  Developmental Toxicity The potential for chlorothalonil to induce developmental toxicity has been investigated in the rat and rabbit. In rabbits, maternal toxicity was evident at 20 mg/kg/d with body weight loss and decreased food consumption observed at this dose (Wilson et al., 1988d). One death occurred in each of the mid and high dose groups. There were no adverse effects on the fetus and no treatment-related effects on the incidence of skeletal or visceral malformations. The NOEL for maternal toxicity was 10 mg/kg/d and the NOEL for developmental toxicity was 20 mg/kg/d. In rats, chlorothalonil was maternally toxic at 400 mg/kg/ d with mortality, decreased body weight gain, and decreased food consumption at this dose (Mizens et al., 1983). Food consumption was significantly decreased at all doses during days 6–9 of gestation and at the top dose on days 9–15. Food consumption returned to normal values on cessation of treatment with a compensatory increase seen at 25 and 100 mg/kg/d. There was a significant increase in the incidence of postimplantation loss due to early embryonic death at 400 mg/kg/d with a corresponding decrease in viable litter size. One rat at this dose level had reabsorbed 16 out of 17 implantation sites. However, exclusion of this animal from the statistical analysis still resulted in a significant increase in postimplantation loss compared to concurrent and historical controls. The NOELs for maternal and developmental toxicity were 100 mg/kg/d. It is concluded that chlorothalonil is not a developmental toxicant when tested up to doses that cause significant maternal toxicity and maternal death.

Chapter | 91  Mammalian Toxicokinetics and Toxicity of Chlorothalonil

Parent Chlorothalonil CN CI

CI

CN

CI CI

Absorption Conjugation to GSH CN

CN

CI

CI

GS

SG

CI

CN

CI

CN

CI

Biliary excretion

Liver GS

Plasma metabolites

GIT

Plasma

β-lyase Kidney

Mercapturates

postpartum at 1500 (68 mg/kg/d) and 3000 ppm (145 mg/ kg/d). This was only seen in the F1b litter at 1500 ppm but was seen consistently across all litters at 3000 ppm. Therefore, NOAEL for fetotoxicity is considered to be 1500 ppm (68 mg/kg/d). The NOEL for reproductive performance was 145 mg/kg/d with no effects at the top dose. It is concluded that chlorothalonil is not a reproductive toxicant. There was no evidence of reproductive toxicity in the absence of maternal toxicity. Therefore, the data are consistent with the view that the fetus and developing animal are not uniquely sensitive to chlorothalonil. Table 91.7 presents a summary of the reproductive toxicity studies. The key NOAELs for all toxicological endpoints are summarized in Table 91.8.

91.10  Investigative Toxicity   Studies

Cysteine conjugates N-acetylase

1961

Thiol-derived conjugates Renal Toxicity

Urinary excretion Figure 91.2  Schematic outlining potential pathways of chlorothalonil metabolism in the rat that lead to formation of toxic metabolites within the kidney. Following absorption from the gastrointestinal tract, chlorothalonil is conjugated to glutathione in the liver. Further metabolic processing results in the formation of cysteine conjugates that may be detoxified via N-acetylase or activated to toxic thiol-derived species. GSH  glutathione, GIT  gastrointestinal tract.

91.9.2  Fertility In a two-generation reproductive toxicity study (Lucas et al., 1990), chlorothalonil caused a dose-related decrease in body weight gain which was evident at all doses in F0 and F1 parental generations although achieving statistical significance at 1500 and 3000 ppm (68 and 145 mg/kg/d). No mortalities or clinical signs of toxicity were observed in this study. Hyperplasia of the forestomach and kidney was observed at all doses in both parental generations with more marked effects in the F1 generation. Thus, a NOEL could not be established for parental toxicity. There were no adverse effects on reproductive performance or development including fertility indices, gestation length, litter size, number of live pups and stillborn pups, and pup survival. No gross malformations were observed which could be considered as treatment-related. There was a significant decrease in mean pup body weight on day 21

91.10.1  Acute Effects on Hepatic and   Renal Glutathione Content This study was designed to investigate and compare the time course effect of the acute oral administration of chlorothalonil on hepatic and renal glutathione (nonprotein sulfhydryl) content (Sadler and Ignatoski, 1985). At 5000 mg/kg chlorothalonil caused an decrease in body weight gain and liver weight which were evident 18 h after treatment. Within 9 h of treatment, hepatic glutathione levels were decreased and renal glutathione levels were elevated. The depletion of hepatic glutathione is considered a direct consequence of glutathione conjugation within the liver utilizing tissue resources. The increase in renal glutathione content is more difficult to explain but may be a consequence of urinary excretion of glutathione conjugates.

91.10.2  Effect of Dietary vs Gavage Dosing on Renal Toxicity in the Rat A study was conducted that was designed to compare the early morphological changes in the rat kidney following oral administration of chlorothalonil by gavage with those following dietary administration (Ford et al., 1988). Chlorothalonil was administered by gavage at 175 and 1750 mg/kg in the diet (equivalent to 88 mg/kg bw). Effects in the kidney were determined at 24, 48, 73, and 96 h postadministration. At the 48 h postadministration time point, vacuolar degeneration of the proximal tubular epithelium was observed in 2/3 animals that were gavaged. After 96 h all animals (gavage and diet) exhibited vacuolar degeneration of proximal tubular epithelium although the incidence of affected tubules was higher in gavaged animals than in those administered chlorothalonil in diet.

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Table 91.7  Summary of Reproductive Toxicity of Chlorothalonil Study

Species

NOAEL (mg/kg bwt/day)

Developmental

Rat

Maternal and developmental toxicity—100 (decreased maternal weight   gain and increased incidence of resorptions at 400 mg/kg/d)

Developmental

Rabbit

Maternal—10 (decreased weight gain at 20 mg/kg/d)   Developmental—20 (top dose)

Two-generation reproduction

Rat

Parental—none (renal and forestomach lesions at all doses; LOAEL was   23 mg/kg/d) Developmental—68 (decreased pup weight at day 21 at 145 mg/kg/d) Reproductive—145 (no effects at top dose)

91.11  Human data Most information concerning the effects of chlorothalonil in humans has been obtained from exposures arising in the manufacture and production and chlorothalonil. Health screening programs in such facilities have shown that the majority of effects documented following exposure to chlorothalonil were attributable to the irritant nature of the substance and included irritation to the skin, eyes, and respiratory tract.

91.11.1  Dermal Effects There are a number of case reports in the published literature documenting occupational dermatitis in workers exposed to technical chlorothalonil. Skin reactions have also been documented in patch test studies with human volunteers. The main criticism of these reports is that they do not clearly discriminate between effects that may be a consequence of skin irritation and those that may represent a true sensitization response. Nevertheless, the weight of evidence suggests that chlorothalonil is a weak skin sensitizer in humans. Special skin surveys were conducted at the main chlorothalonil manufacturing plant to compare dermal findings in 1978 and 1979. In 1978, 60% of the employees had some type of skin abnormality including 19 cases of contact dermatitis (McAmis, 1994a). In 1979, following the initiation of improved industrial hygiene measures in late 1978, there were no cases of contact dermatitis and only 21% of the workers had some kind of skin abnormality. The most common abnormality was skin drying which was seen in 19 of the 26 employees with skin abnormalities (Chelsky, 1980a, b). A delayed irritant dermatitis has been documented which may occur up to 72 h after ­ exposure

and, although photosensitization reactions may occur, they are very rare events.

91.11.2  Ocular Effects A review of clinical cases from employees exposed to chlorothalonil, at a packaging plant where the exposure was described as infrequent and light, was conducted in 1990 (Chelsky, 1990a). The purpose of this review was to assess the effects of chlorothalonil on the human eye. All of the ocular exposures to chlorothalonil involved intense pain with mild to moderate conjunctivitis and irritation of the corneal surface. Ocular edema was also seen in more extensive exposures. With lesser exposures, complete recovery occurred within 24 h. Recovery took slightly longer with following extensive exposure. In no instance was corneal opacity observed.

91.11.3  Respiratory Effects Where respiratory effects have been noted these are generally consistent with the irritant properties noted in animal studies, although of a much less severe nature. In a review of medical records from workers at an independent facility used to grind technical chlorothalonil (Chelsky, 1990b, 1992; McAmis, 1994b), it was noted that, even in a workplace described as “dusty” and with workers wearing little protective clothing, ocular and dermal effects predominated, although a few cases of nasal and pharyngeal pain, burning, and soreness were noted. In a study conducted at another manufacturing facility, workers exposed to chlorothalonil showed a lower forced expiratory volume and higher incidences of nose and throat irritation, coughing phlegm, and shortness of breath than reference workers (Huang et al., 1995). Thus, the

Chapter | 91  Mammalian Toxicokinetics and Toxicity of Chlorothalonil

1963

Table 91.8  Summary of Key Toxicological Endpoints Study

Endpoint

NOEL (mg/kg/d)

LOEL (mg/kg/d)

90 day rat diet

Increased kidney weight

1.5

3

Hyperplasia of forestomach

3

10

Renal hyperplasia

10

40

Increased kidney weight

3

9

Hyperplasia of forestomach

3

9

Renal hyperplasia

48

130

90 day mouse diet

90 day dog (capsule)

Decreased body weight gain

  15

150

1 year dog (capsule)

Decreased body weight gain

150

500

2 year rat diet

Renal hyperplasia Renal tumors Forestomach hyperplasia Forestomach tumors

1.8 3.8 1.8 1.8

3.8 15 3.8 3.8

2 year mouse oral

Renal hyperplasia Renal tumors Forestomach hyperplasia Forestomach tumors (squamous cell)

5.4 99 (top dose) 1.6 1.6

23 None 5.4 5.4

21 day dermal—rat

Local effects in skin Systemic toxicity

None None–but NOEL for   renal hyperplasia of    600

60 (lower dose)   60 (23% decrease in body    weight gain at lowest     dose)

Rat reproductive toxicity

Parental NOEL (renal and forestomach hyperplasia) None   68 Developmental (decreased pup weight at day 21)

Developmental toxicity

Rat: Maternal (decreased weight gain) Developmental (increased resorptions) Rabbit: Maternal (decreased weight gain) Developmental

i­nformation available from animal and human exposure indicates that chlorothalonil is irritating to the respiratory tract; a finding that is entirely consistent with the local site-ofcontact toxicity seen in other epithelial tissues.

91.11.4  Clinical Cases and Poisoning Incidents Considering that chlorothalonil has been a commercial fungicide for over 25 years, there have been few reports of adverse effects in humans resulting from its use. The majority of the reported human effects have been related

23 (lower dose) 145

100 100

400 400

  10   20

20 None

to the irritant properties of chlorothalonil. Of the reported skin effects, contact dermatitis is the most frequent diagnosis and this finding is almost exclusively in individuals exposed to chlorothalonil for prolonged periods (over 8 h) in an occupational environment. In summary, there have been a few reports in the literature of humans suffering adverse health effects following exposure to chlorothalonil (McAmis, 1995). Considering that chlorothalonil has been marketed for more than 25 years as a fungicide in agriculture, forestry, nursery plants, paints, and stains, the reports of adverse effects are very rare. The reported effects are associated with the irritant properties of the technical material.

1964

References Andre, J. C., et al. (1991a). “Comparison of the Effects of Dose Level and Vehicle on the Dermal Absorption of 14CChlorothalonil by Male Rats.” Unpublished Syngenta study, Rep. 1698-88-0007-AM-001. Andre, J. C., et al. (1991b). “Evaluation of Mitochondrial Function in the Presence and Absence of Sulfur-Containing Analogs of Chlorothalonil.” Unpublished Syngenta study, Rep. 3113-88-0107-AM-001. Auletta, A., and Kouri, R. (1977). “Activity of DTX-77-0033 in a Test for differential Inhibition of Repair Deficient and Repair Competent Strains of Salmonella typhimurium.” Unpublished Syngenta study, Rep. 000-5TX-77-0033-002. Banzer, C. B., and Kouri, R. E. (1977). “Activity of Chlorothalonil in the Salmonella Microsomal Assay for Bacterial Mutagenicity.” Unpublished Syngenta study, Rep. 000-5TX-77-0035-001. Chelsky, M. (1980a). “Special Skin Surveys of Green’s Bayou Plant Employees, 1978 and 1979.” Confidential company medical report originally generated by Diamond Shamrock Corp. Chelsky, M. (1980b). “Skin rashes Among Green’s Bayou Plant Employees.” Confidential company medical report originally generated by Diamond Shamrock Corp. Chelsky, M. (1990a). “Study of Chlorothalonil Plant Workers 1990. Evaluation of Potential for Persistent Effects on Eyes of Workers.” Confidential company medical report originally generated by Diamond Shamrock Corp. Chelsky, M. (1990b). “Annual Employee Health Screening Greens Bayou Plant, 1986–1990. Special Reference to Chlorothalonil Workers and the Respiratory System.” Confidential company medical report originally generated by Diamond Shamrock Corp. Chelsky, M. (1992). “Annual Employee Health Screening Reports, Greens Bayou Plant, 1986–1991. Special Reference to Chlorothalonil Workers and the Respiratory System.” Confidential company medical report originally generated by Diamond Shamrock Corp. Fillmore, G., et al. (1993). “A 90-Day Oral Toxicity Study in Dogs with Chlorothalonil.” Unpublished Syngenta study, Rep. 5210-920103-TX-003. Ford, W. H., et al. (1988). A 4-Day Study in Rats with Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 1095-86-0091TX-002. Francis et al. (1973). “Daconil Technical Air Milled, Eye Irritation in the Albino Rabbit.” Unpublished Syngenta study, Rep. 7948-95-3. Hironaka, M., et al. (1996). “Analysis of Hyperplastic Changes in the Stomach and Kidney of Male Rats after 28-day Induction by Chlorothalonil Technical.” Unpublished Syngenta study, Rep. 3561. Huang, J. et al. (1995). Respiratory effects and skin allergy in workers exposed to tetrachloroisophthalonitrile. Bull. Environ. Contam. Toxicol. 55, 320–324. Jones, R. E., et al. (1984). “Salmonella/Mammalian-Microsorne Plate Incorporation Assay (Ames Test) with and without Renal Activation with Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 694-5TX-84-0064-002. Killeen, J. C., Jr. (1983). “Research on the Possible Mutagenic Potentiality of Chlorothalonil by the Detection of Chromosomal Alteration in the Rat (Rat, Mouse and Hamster).” Unpublished Syngenta study, Rep. 000-5TX-81-0025-001. Killeen, J. C., Jr., and Siou, G. (1983). “The Micronucleus Test in the Rat, Mouse and Hamster Using Chlorothalonil.” Unpublished Syngenta study, Rep. 000-5TX-81-0024-004. Kouri (1977).

Hayes’ Handbook of Pesticide Toxicology

Lucas, F., et al. (1990). “A Two Generation Reproduction Study in Rats with Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 1722-87-0121-TX-003. Magee, T. A., et al. (1990). “Study to Evaluate the Urinary Metabolites of Chlorothalonil Following Dermal Application to Male Rhesus Monkeys.” Unpublished Syngenta study, Rep. 3382-89-0214-AM-001. Marciniszyn, J. P., et al. (1985a). “Pilot Study of the Biliary Excretion of Radioactivity Following Oral Administration of Chlorothalonil (14CDS-2787) to Sprague–Dawley Rats.” Unpublished Syngenta study, Rep. 633-4AM-83-0062-002. Marciniszyn, J. P., et al. (1985b). “Study of the Distribution of Radioactivity Following Oral Administration of 14C-Chlorothalonil (14C-SDS-2787).” Unpublished Syngenta study, Rep. 631-4AM-84-0078-002. Marciniszyn, J. P., et al. (1986a). “Study of the Biliary Excretion of Radioactivity Following Oral Administration of 14C-Chlorothalonil (14C-DS-2787) to Male Sprague–Dawley Rats.” Unpublished Syngenta study, Rep. 633-4AM-85-0012-002. Marciniszyn, J. P., et al. (1986b). “Pilot Study of the Effect of the Gamma-Glutamyl Transpeptidase Inhibitor, AT-125 on the Metabolism of 14C-Chlorothalonil.” Unpublished Syngenta study, Rep. 1376-86-0072-AM-002. McAmis, R. J. (1994a). “Review of Dermal Chlorothalonil Exposures in Humans.” Confidential company medical report originally generated by Diamond Shamrock Corp. McAmis, R. J. (1994b). “Review of Respiratory Chlorothalonil Exposures in Humans.” Confidential company medical report originally generated by Diamond Shamrock Corp. McAmis, R. J. (1995). “Diagnosis of Poisoning, Specific Signs of Poisoning, Clinical Tests.” Confidential company medical report originally generated by Diamond Shamrock Corp. Mead, R. L., et al. (1987a). “Analysis of Urine Samples from a 90-Day Feeding Yes No Study in Rats with the Monoglutathione Conjugate of Chlorothalonil (T-117-11).” Unpublished Syngenta study, Rep. 1108-85-0078-TX-006. Mead, R. L., et al. (1987b). “Analysis of Urine Samples from a 90-Day Feeding Study in Rats with Chlorothalonil (T-117-11).” Unpublished Syngenta study, Rep. 1115-85-0079-TX-005. Mizens, M., and Laveglia (1994). “A Chronic (12-Month) Oral Toxicity Study in Dogs with Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 92-0457. Mizens, M., et al. (1983). “A Teratology Study in Rats with Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 517-5TX-82-0011-003. Mizens, M., et al. (1986a). “A 21-Day Repeated Dose Dermal Toxicity Study Rats with Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 68-59-96-0113-TX-02. Mizens, M., et al. (1986b). “In Vitro Chromosomal Aberration Assay in Chinese Hamster Ovary (CHO) Cells with Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 1109-85-0082-TX-002. Mizens, M., et al. (1996). “A 90-Day Pilot Study for the Evaluation Proliferation in the Kidneys of Male Rats Following the Oral Administration of Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 6704-96-0010-TX-003. O’Meara, H. O., and Laveglia, J. (1995). “Eye Irritation Study in Albino Rabbits with Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 6300-95-0083-TX-001. Moore (2000). Price, P., and Ballee, D. (1979). “Analyses of Samples from Cell Transformation Studies for 2,4,5,6-Tetrachloroisophthalonitrile

Chapter | 91  Mammalian Toxicokinetics and Toxicity of Chlorothalonil

(Chlorothalonil, DS-2787) and 4-Hydroxy-2,5,6-TrichloroisoPhthalonitrile (DS-3701) (DTX-77-0037 and DTX-77-0041).” Unpublished Syngenta study, Rep. 041-5TX-79-0021-001. Proudlock, R. J. (1995). “In Vivo Bone Marrow Chromosomal Analysis in Chinese Hamsters Following Multiple Dose Administration of Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 600594-0047-TX-003. Ribovich, M. L., et al. (1982). “Balance Study of the Distribution of Radioactivity Following Oral Administration of 14C-Chlorothalonil (14C-DS-2787) to Male Mice.” Unpublished Syngenta study, Rep. 613-4AM-82-0178-001. Sadler, E. M., and Ignatoski, J. A. (1985). “Time Course of the Acute Effect of Technical Chlorothalonil on Hepatic and Renal Glutathione Content in Male Rats.” Unpublished Syngenta study, Rep. 751-5TX-85-0032-001. Savides, M. C., et al. (1985). “Pilot Study for the Determination of the Effects of Probenecid Pre-treatment on Urinary Metabolites and Excretion of 14C-Chlorothalonil (14C-SDS-2787) Following Oral Administration to Male Sprague–Dawley Rats.” Unpublished Syngenta study, Rep. 621-4 AM-85-0035-001. Savides, M. C., et al. (1986a). “Study of the Distribution of Radioactivity Following Repeated Oral Administration of 14C-Chlorothalonil to Male Sprague–Dawley Rats.” Unpublished Syngenta study, Rep. 1173-84-0079-AM-003. Savides, M. C., et al. (1986b). “Identification of Metabolites in Urine and Blood Following Oral Administration of 14C-Chlorothalonil to Male Rats: Effects of Multiple Dose Administration on the Excretion of Thiol Metabolites in Urine.” Unpublished Syngenta study, Rep. 621-4AM-83-0061-002. Savides, M. C., et al. (1987). “Determination of the Covalent Binding of Radio-label to DNA in the Kidneys of Male Rats Administered 14CChlorothalonil (14C-SDS-2787).” Unpublished Syngenta study, Rep. 1173-86-0096-AM-002. Savides, M. C., et al. (1988). “A Study to Evaluate the Effects of SulfurContaining Analogs of Chlorothalonil on Mitochondrial Function.” Unpublished Syngenta study, Rep. 1479-87-0037-AM-001. Savides, M. C., et al. (1989). “Study to Determine the Metabolic Pathway for Chlorothalonil Following Dermal Application to Rats.” Unpublished Syngenta study, Rep. 1625-87-0057-AM-001. Savides, M. C., et al. (1990). “Study to Evaluate the Urinary Metabolites of Chlorothalonil from Male Rhesus Monkeys.” Unpublished Syngenta study, Rep. 3349-89-0179-AM-001. Savides, M. C., et al. (1995). “Study to Determine the Extent and Nature of Yes No Biliary Excretion of Chlorothalonil and/or Metabolites in the Dog.” Unpublished Syngenta study, Rep. 5521-93-0319-AM-001. Shirasu, Y. (1978). “Mutagenicity Testing on Daconil in Microbial Systems.” Unpublished Syngenta study, Rep. 000-5TX-61-0002-001. Shults and Wilson (1985). “Dermal Sensitisation Study (Closed Patch Repeated Insult) in Guinea Pigs with Chlorothalonil 90DG Formulation.” Unpublished Syngenta study, Rep. 707-5TX-840126-002. Shults, S. K., et al. (1981a). “Acute Oral Toxicity (LD50) Study in Rats with Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 296-5TX-80-0092-002. Shults, S. K., et al. (1981b). “Acute Dermal Toxicity (LD50) Study in Albino Rabbits with Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 296-5TX-80-0093-002. Shults, S. K., et al. (1981c). “Acute Inhalation Toxicity Study (Four Hour Exposure) in Rats with Technical Chlorothalonil (SDS-2787).” Unpublished Syngenta study, Rep. 296-5TX-80-0096-002.

1965

Shults, S. K., et al. (1981d). “Primary Dermal Irritation Study in Albino Rabbits with Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 296-5TX-80-0094-002. Shults, S. K., et al. (1983). “A 90 Day Feeding Study in Mice with 2,4,5,6-Tetrachloroisophthalonitrile (Chlorothalonil).” Unpublished Syngenta study, Rep. 618-5TX-83-0007-004. Shults, S. K., et al. (1985). “Histopathologic Re-evaluation of Renal Tissue from a 90-Day Feeding Study in Mice with Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 753-5TX-85-0053-002. Shults, S. K., et al. (1986). “21-Day Repeated Dose Dermal Toxicity Study in Albino Rabbits with Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 754-5TX-85-0023-007. Shults, S. K., et al. (1993). “Acute (Four-Hour) Inhalation Toxicity (LC50) Study in Rats with Hammer-Milled Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 5290-92-0160-TX-002. Siou, G., et al. (1985). “Acute and Subchronic In Vivo Bone Marrow Chromosomal Aberration Assay in Chinese Hamsters with Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 625-5TX-83-0014-003. Tucker, S. B. (1986). “Skin Sensitisation Studies with Chlorothalonil Conducted at the Department of Occupational Dermatology, University of Texas.” Unpublished Syngenta studies, Reps. 5TX-840023, 5TX-84-0027, 5TX-84-0012, 1094-84-0012-DA002, 5TX-840044, 5TX-84-0076, and 5TX-84-0045. Ward, R. J. (1989a). “Chlorothalonil: In Vitro Absorption from Technical Material through Human Epidermis.” Unpublished Syngenta study, Central Toxicology Laboratory, Rep. CTL/P/2640. Ward, R. J. (1989b). “Chlorothalonil: In Vitro Absorption from Bravo 720 Formulation through Human Epidermis.” Unpublished Syngenta study, Central Toxicology Laboratory, Rep. CTL/P/2880. Wazeter and Lucas (1971). “Acute Intraperitoneal Toxicity (LD50) in Male Albino Rats of Technical Chlorothalonil.” Unpublished Syngenta studies, Rep. 000-5TX-71-0006-001. Wilkinson, C. F., and Killeen, J. C. (1996). A mechanistic interpretation of the oncogenicity of chlorothalonil in rodents and an assessment of human relevance. Regulatory Toxicol. Pharmacol. 24, 69–84. Wilson, P. D. (1977a). “Primary Eye Irritation Study in Rabbits.” Unpublished Syngenta study, Rep. DTX-77-0075. Wilson, P. D. (1977b). “Primary Eye Irritation Study in Rabbits.” Unpublished Syngenta study, Rep. DTX-77-0059. Wilson, P. D. (1977c). “Primary Eye Irritation Study in Rabbits.” Unpublished Syngenta study, Rep. DTX-77-0069. Wilson, N. H., et al. (1982a). “Dermal Sensitisation Study in HartleyDerived Guinea Pigs with Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 394-5TX-81-0132-002/7020. Wilson, N. H., et al. (1982b). “Four Week Dietary Range-Finding Study in Rats with Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 099-5TX-81-0174-003. Wilson, N. H., et al. (1983a). “A 90-Day Toxicity Study of Technical Chlorothalonil in Rats.” Unpublished Syngenta study, Rep. 099-5TX-80-0200-006. Wilson, N. H., et al. (1983b). “A Subchronic Toxicity Study of Technical Chlorothalonil in Rats.” Unpublished Syngenta study, Rep. 562-5TX-81-0213-004. Wilson, N. H., et al. (1983c). “A Chronic Dietary Study in Mice with Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 108-5TX-79-0102-004. Wilson, N. H., et al. (1984). “A Subchronic Toxicity Study of Technical Chlorothalonil in Rats (Electron Light Microscopy of Kidneys).” Unpublished Syngenta study, Rep. 562-5TX-8-0213-004-001.

1966

Wilson, N. H., et al. (1985a). “Histopathologic Re-evaluation of Renal Tissue from a 90-Day Toxicity Study in Rats with Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 753-5TX-85-0055-002. Wilson, N. H., et al. (1985b). “Histopathologic Re-evaluation of Renal Tissue from a Subchronic Toxicity Study of Technical Chlorothalonil in Rats.” Unpublished Syngenta study, Rep. 753-5TX-85-0056-002. Wilson, N. H., et al. (1985c). “A Tumourgenicity Study of Technical Chlorothalonil in Rats.” Unpublished Syngenta study, Rep. 099-5TX-80-0234-008. Wilson, N. H., et al. (1987). “A Tumourgenicity Study of Technical Chlorothalonil in Male Mice.” Unpublished Syngenta study, Rep. 1099-84-0077-TX-006. Wilson, N. H. (1988a). “Guinea Pig Maximization Test with Technical Chlorothalonil (T-117-11).” Unpublished Syngenta study, Rep. 1094-84-0044-TX-001.

Hayes’ Handbook of Pesticide Toxicology

Wilson, N. H. (1988b). “Guinea Pig Epicutaneous Test Involving Chlorothalonil in Acetone.” Unpublished Syngenta study, Rep. 1094-84-0045-TX-001. Wilson, N. H. (1988c). “Guinea Pig Maximization Test with Technical Chlorothalonil (T-117-11).” Unpublished Syngenta study, Rep. 1094-84-0076-TX-001. Wilson, N. H., et al. (1988d). “A Teratology Study in Rabbits with Technical Chlorothalonil.” Unpublished Syngenta study, Rep. 1544-87-0060-TX-002. Wilson, N. H., et al. (1989). “A Tumourgenicity Study of Technical Chlorothalonil in Rats.” Unpublished Syngenta study, Rep. 1102-84-0103-TX-007.

Chapter 92

Inhibitors of Aromatic Acid Biosynthesis Donna Farmer Monsanto Company

92.1  Introduction

92.2  Glyphosate

Glyphosate is a broad-spectrum, postemergent systemic herbicide with activity on essentially all annual and perennial plants. Glyphosate-based formulations are used worldwide in virtually every phase of agricultural, industrial, silvicultural, and residential weed control. Due to low solubility in water, glyphosate is typically formulated into commercial products in the form of a salt. Glyphosate is poorly absorbed both dermally and via oral exposure and it is not biotransformed. It has been shown that glyphosate does not bioaccumulate. Animal studies indicate that glyphosate is essentially nontoxic via acute oral and dermal exposure, and that glyphosate salts are non-irritating to the eyes and skin. Glyphosate does not produce dermal sensitization in guinea pigs. In repeated dose studies in laboratory animals, treatment-related effects included reduced body weight gain, increased liver weights, degenerative ocular lens changes, and microscopic liver changes but only at very high dose levels (approximately 2–3% of the diet). No treatment-related tumors have been found in multiple carcinogenicity studies. Glyphosate has consistently produced negative results in standard mutagenicity assays conducted according to international guidelines. Regulatory agencies and other scientific organizations have concluded that glyphosate is neither carcinogenic nor mutagenic. The U.S. Environmental Protection Agency (U.S. EPA) USEPA has classified glyphosate in Category E (“Evidence of Non-Carcinogenicity in Humans”). There is no evidence of developmental or reproductive effects resulting from glyphosate exposure. In humans, accidental exposure to glyphosate formulations may result in minor, transient ocular and dermal irritation, but serious effects have not been observed.

92.2.1  Identity, Properties, and Uses

Hayes’ Handbook of Pesticide Toxicology Copyright © 2001 Elsevier Inc. All rights reserved

Chemical Name  Glyphosate is N-(phosphonomethyl) glycine. Structure  The structure of glyphosate is shown in Fig. 92.1. Synonyms  The common name glyphosate is in general use. Trade names include Roundup®, RoundupUltra®, Roundup-Pro®, Landmaster®, Rodeo®, Accord®, Spark®, Vision®, and Biactive®. The CAS registry number for the acid is 1071-83-6. Physical and Chemical Properties  Glyphosate acid is typically referred to as the technical grade material and has the empirical formula C3H8NO5P. It is a white, odorless, crystalline powder with a melting point of 184.5°C, a molecular weight of 169.1, and a specific gravity of 1.704. Glyphosate is not flammable, is not explosive and has a vapor pressure of 1.84  107 mm Hg at 45°C. Glyphosate is a relatively strong acid with a pH of 2 in 1% aqueous solution. The solubility of glyphosate in water is 1.2 wt% at 25°C and approximately 6 wt% at 100°C. It is slightly soluble in a few strong organic acids but relatively insoluble in most organic solvents. Because of its limited solubility in water, commercial herbicide formulations contain

O

O H N

C HO

C H2

OH

P C H2

OH

Figure 92.1  Glyphosate acid.

1967

1968

glyphosate in the form of a salt (i.e., isopropylamine, ammonium, phosphonium, etc.) (Franz et al., 1997). History, Formulations, and Uses  Gyphosate is a broad-spectrum, nonselective, postemergent, systemic herbicide with activity on essentially all annual and perennial plants. The herbicidal properties of glyphosate were discovered by Monsanto in 1970, and the first commercial formulations were introduced in 1974 under the Roundup brand name. Today, glyphosate-based formulations are used in over 100 countries in virtually every phase of agricultural, industrial, silvicultural, and residential weed control, making it one of the most important weed-pest control tools ever introduced. Agricultural use of glyphosate continues to expand. It has contributed significantly to the growing worldwide adoption of conservation and reduced tillage techniques as well as applications involving genetically modified plant varieties which can tolerate glyphosate treatment.

92.2.2  Toxicity to Laboratory Animals Standard toxicity studies have been performed with technical grade glyphosate (averaging 96% purity on a dry weight basis). The results have been summarized in the reregistration eligibility decision (RED) document issued by the United States Environmental Protection Agency (U.S. EPA, 1993), the World Health Organization (WHO, 1994), and Williams et al. (2000). These results demonstrate that glyphosate has very low acute toxicity and is not mutagenic, teratogenic, carcinogenic, or a reproductive toxicant. Acute Studies  The oral LD50 in rats is 5000 mg/kg and the dermal LD50 in rabbits is 5000 mg/kg (WHO, 1994). An acute rat inhalation study has not been conducted with glyphosate technical because it is a nonvolatile solid material which would not generate a respirable vapor or particulate under circumstances of normal use. However, a 4-hour LC50 for an aqueous solution of the isoproplyamine salt of glyphosate was shown to be greater than the highest attainable atmospheric concentration of 1.3 mg/l (Dudek, 1987). Irritation and Sensitization Studies  Glyphosate technical produced mild skin irritation after a single 4-hour exposure. However, glyphosate did not produce dermal sensitization in guinea pigs (U.S. EPA, 1993). Glyphosate, when applied undiluted and without a wash, was severely irritating to the eyes (WHO, 1994). In contrast, the neutral pH isoproplyamine and monosodium salts are nonirritating to the eyes (Branch, 1981; Busch, 1987). Dose Studies  Several subchronic and chronic toxicology studies have been conducted on glyphosate, and the results of these investigations have been reported by the U.S. EPA (1993), WHO (1994), and Williams et al. (2000). The major findings of these studies are summarized in this chapter. In a 3-month feeding study with Sprague-Dawley rats, the no-observed-effect level (NOEL) was 20,000 ppm

Hayes’ Handbook of Pesticide Toxicology

(approximately 1445 mg/kg/day), the highest dose tested. Administration of glyphosate to CD-1 mice for 3 months at dietary levels of 0, 5000, 10,000, and 50,000 ppm resulted in reduced body weight gains in high-dose animals. The NOEL was 10,000 ppm (approximately 2300 mg/kg/day). Glyphosate was applied to the shaven intact and abraded skin of New Zealand white rabbits for 6 hour per day, 5 days per week for 3 weeks at dose levels of 0, 100, 1000, and 5000 mg/kg/day. A slight degree of dermal irritation was observed at the site of application in the high-dose group. No adverse effects were noted in the hematologic, biochemical, and histopathological evaluations. The systemic NOEL was considered to be 5000 mg/kg/day. Glyphosate was given to beagle dogs via oral capsule at dosages of 0, 20, 100, or 500 mg/kg/day for 1 year. No treatmentrelated effects were noted even at the highest dose tested; therefore, the NOEL was considered to be 500 mg/kg/day. Brahman-cross heifers received daily dosages of the isopropylamine salt of glyphosate via stomach tube for seven consecutive days at dosages of 0, 540, 830, 1290, and 2000 mg/kg/day. Mortality was observed only at the two highest doses. Other effects, including body weight loss, diarrhea, serum chemistry changes, and histopathological findings were observed at or above 830 mg/kg/day. Changes in several hematologic parameters observed at 1290 mg/ kg/day and above were considered secondary to fluid and blood volume alterations resulting from the diarrhea. The NOEL was considered to be 540 mg/kg/day. Three rodent bioassays were conducted with glyphosate. In the first of two long-term feeding studies conducted in SpragueDawley rats, glyphosate was administered in the diet at concentrations of 0, 60, 200, and 600 ppm for approximately 26 months. The NOEL was considered to be 600 ppm (32 mg/kg/day) because no tumors or other adverse effects related to treatment were noted at any dose level. In the second chronic study, rats were fed glyphosate in the diet at concentrations of 0, 2000, 8000, and 20,000 ppm for approximately 2 years. No tumors related to treatment were observed. The only effects considered related to treatment were observed in high-dose animals and included decreased body weight gain in females and degenerative ocular lens changes, increased liver weights, and elevated urine pH or specific gravity in males. The NOEL in this study was concluded to be 8000 ppm (409 mg/kg/day). Glyphosate was fed to CD-1 mice in the diet at concentrations of 0, 100, 5000, or 30,000 ppm. No treatment-related tumors were observed. The NOEL in this study was concluded to be 5000 ppm (750 mg/kg/day) based upon reduced body weight gains in high-dose males and females and microscopic liver changes (central lobular hepatocyte hypertrophy and hepatocyte necrosis) in high-dose males. Absorption, Distribution, Metabolism, and Excretion  Absorption of glyphosate across skin and gastrointestinal membranes is minimal. In vitro absorption of glyphosate through human skin was no more than 2% of

Chapter | 92  Inhibitors of Aromatic Acid Biosynthesis

applied dose (Wester et al., 1991). Wester et al. (1991) also reported the in vivo dermal absorption of glyphosate in the rhesus monkey to be 2.2% at a high dose of 5400  g/cm2. The results of several studies show that there is rapid elimination, no biotransformation, and minimal tissue retention of glyphosate in various species, including mammals, birds, and fish (U.S. EPA, 1993; WHO, 1994). Greater than 90% of an orally administered dose of glyphosate is rapidly eliminated in 72 hours (National Toxicology Program, 1992). Typically, approximately 70% of the administered dose is eliminated in the feces, with the remainder eliminated in the urine. In all cases, less than 0.5% of the administered dose is found in the tissues and organs, demonstrating that glyphosate does not bioaccumulate in edible tissues. Studies of the metabolism of glyphosate in experimental animals (rats, rabbits, lactating goats, and chickens) indicate that it is not biotransformed, with essentially all the administered dose excreted as unchanged parent molecule (Bodden, 1988; Colvin and Miller, 1973; Ridley and Mirly, 1988). Genotoxicity Studies  Glyphosate was negative in well-validated mutagenicity assays performed for regulatory purposes conducted according to international guidelines under good laboratory practices (U.S. EPA, 1993; WHO, 1994). These assays assessed a variety of end points both in vitro and in vivo and included the following: Salmonella typhimurium (Ames assay), Escherichia Coli WP-2 reverse mutation, rec-assay with Bacillus subtilis, CHO/HGPRT, in vivo mouse bone marrow micronucleus, and in vitro hepatocyte primary culture-DNA repair assay. Williams et al. (2000), in a review on glyphosate, employed a weight-of-evidence evaluation of the many genotoxicity assays including those submitted for regulatory purposes as well as others in the published scientific literature. It was concluded that glyphosate is neither mutagenic nor clastogenic. A limited number of studies in the literature have reported positive results regarding the genotoxic potential of glyphosate; review by these authors found that these assays used toxic dose levels, irrelevant routes of exposure, end points, and test systems, and/or deficient testing methodology. Carcinogenicity Studies  Regulatory agencies and other scientific organizations have concluded that glyphosate is neither carcinogenic nor mutagenic. In June of 1991, the U.S. EPA following a thorough review of all toxicology data available concluded that glyphosate should be classified in Category E (“Evidence of Non-Carcinogenicity in Humans”). This classification was based upon the observation of no treatment-related tumors at any dose level with glyphosate tested up to the limit dose in rats and up to levels higher than the limit dose in mice, and upon the lack of evidence for mutagenicity with glyphosate (U.S. EPA, 1992). Mode of Action  Glyphosate’s mode of action has been previously described in detail (Franz et al., 1997). Glyphosate

1969

inhibits plant growth through competitive ­ inhibition of the enzyme 5-enolpyruvoylshikimate 3-­phosphate synthase (EPSPS). This enzyme plays a key role in the biosynthesis of the intermediate, chorismate, necessary for the ­ synthesis of the essential amino acids phenylalanine, tyrosine, and tryptophan. This aromatic amino acid biosynthetic pathway (shikimic acid pathway) is found in plants as well as some fungi and bacteria but not in insects, birds, fish, mammals, and humans, thus providing a specific selective toxicity to plant species. Developmental and Reproduction Studies  The reproductive and developmental toxicity data base has been evaluated by the U.S. EPA (1999). This assessment was conducted under the Federal Food, Drug, and Cosmetic Act, as amended by the Food Quality Protection Act (FQPA). The FQPA was enacted by the U.S. Congress in 1996 with provisions that support the governments focus on children’s environmental health risks. It requires the U.S. EPA to more carefully consider the risks posed to infants and children by pesticide residues on food when setting acceptable residue levels and tolerances. The U.S. EPA is required to apply an additional 10-fold safety factor to ensure the protection of infants and children unless a determination can be made on the basis of reliable data that a lesser margin of safety is protective. As a result of their assessment, the EPA may also require additional testing to detect potential developmental neurotoxic effects. Regarding glyphosate, EPA has concluded that there is a complete toxicity data base and exposure data is complete or can be estimated based on data that reasonably accounts for potential exposures. It was concluded there is no indication that the developing fetus or neonate is more sensitive than adult animals. Consequently no developmental neurotoxicity studies were required. The EPA believes that reliable data support the use of the standard 100-fold uncertainty factor and concluded that there is a reasonable certainty that no harm will result to infants and children from aggregate exposure to glyphosate residues. The studies supporting these conclusions are summarized next. Sprague-Dawley rats were dosed by gavage at doses of 0, 300, 1000, or 3500 mg/kg/day during days 6–19 of gestation. At 3500 mg/kg/day, the following signs of toxicity were observed: increased mortality (6 of 25 dams died) and other clinical signs of toxicity, decreased fetal weights, increased incidence of early resorptions, decreases in total number of implantations and the number of viable fetuses, and increased number of fetuses with reduced ossification of sternebrae. At the lower dose levels these effects were absent. There was no evidence of teratogenicity at any dose level. The NOEL for both maternal and developmental toxicity was 1000 mg/kg/day. In Dutch belted rabbits, glyphosate was tested at dose levels of 0, 75, 175, or 350 mg/kg/day from days 6 through 27 of gestation. The maternal NOEL was determined to be 175 mg/kg/day based on maternal signs of toxicity seen only at the highest dose

1970

tested. These effects included death (10 of 16 does died), diarrhea, and nasal discharge. Excessive maternal mortality resulted in an inadequate number of fetuses for evaluation at 350 mg/kg/day. Therefore, although no developmental toxicity was observed at any dose level, the developmental NOEL was considered to be 175 mg/kg/day. Glyphosate was administered to Sprague-Dawley rats in the diet at dosages of 3, 10, and 30 mg/kg/day for three successive generations (2 litters per generation). There were no treatment-related effects on mating, fertility, or other reproductive parameters. An equivocal increase was noted for the incidence of unilateral renal tubular dilation in male pups of the F3b generation in the high-dose group. The small increase in incidence was not considered related to treatment. This conclusion is supported by the absence of a similar effect in a more recent study which evaluated substantially more animals and used significantly higher dose levels (3% of the diet). In the more recent reproduction study, Sprague-Dawley rats were administered glyphosate in the diet at dosages of 0, 2000, 10,000, and 30,000 ppm (equivalent to 0, 100, 500, and 1500 mg/kg/ day). There was no effect on the ability of treated rats to mate, conceive, carry, or deliver normal offspring. The systemic NOEL was 10,000 ppm based on soft stools and decreased body weights and pup weights during the second and third weeks of lactation. The reproductive NOEL was the highest dose tested, 30,000 ppm.

92.2.3  Human Experience No evidence was observed for the induction of photoirritation nor of allergic or photoallergic contact dermatitis when Roundup herbicide (41% IPA salt of glyphosate, water, and a surfactant) was evaluated in 346 volunteers (Maibach, 1986). Roundup was less irritating than a standard dishwashing detergent and a general all-purpose cleaner and was no different than baby shampoo. Acquavella et al. (1999) evaluated effects from 1513 human ocular exposures to various Roundup formulations reported to an American Association of Poison Control Centers (AAPCC) certified regional poison control center during the years 1993 through 1997. The majority of the reported exposures were judged by the poison center specialists to result in either no injury (21%) or transient minor symptoms (70%). In no case did exposure result in permanent change to the structure or function of the eye. The exposure potential of the general population and applicators to glyphosate have been reviewed by Williams et al. (2000). Exposure of the general population to glyphosate is very low and occurs primarily from the diet. Glyphosate has been registered for use in food crops for over 20 years, and glyphosate is now used in a wide range of crops. The initial uses for glyphosate were for preplanting or preemergence applications and resulted in negligible residues in the crops. Later uses have included applications when the

Hayes’ Handbook of Pesticide Toxicology

crops are present, either using directed spray techniques, applications close to harvest, or herbicide-tolerant crops. These uses can result in residues in edible commodities, although they are still at very low levels. The reference dose (RfD) for glyphosate based on the developmental toxicity study with rabbits (NOEL of 175 mg per kilogram of body weight per day) and using a 100-fold safety factor is calculated to be 2.0 mg(kg body weight)day (U.S. EPA, 1999). The RfD represents the level at or below which daily aggregate dietary exposure over a lifetime will not pose appreciable risks to human health. The U.S. EPA generally has no concern for exposures below 100% of the RfD. The theoretical maximum residue contributions (TMRC) and percentage of RfDs for the U.S. population was estimated to be 0.029960 or 1.5% of the RfD and 0.064388 or 3.2% or the RfD for children (1–6 years old) (U.S. EPA, 1999). Because the qualitative nature of glyphosate residues is well understood and the aggregate exposure is not expected to exceed 100% of the RfD, the U.S. EPA concludes that there is reasonable certainty that no harm will result from aggregate exposure to glyphosate residues. Dermal contact is the most likely route of exposure for applicators; activities such as mixing and loading of glyphosate and extended applications using hand sprayers have the highest potential for exposure. Inhalation is considered to be a minimal route of exposure under most circumstances because of glyphosate’s extremely low vapor pressure. Biological measurements estimating the amount of pesticide that has penetrated into the body, the internal dose, provide the most relevant information for safety assessments. Lavy et al. (1992) found that, of 355 daily urine samples analyzed from silvicultural workers, none contained quantifiable levels of glyphosate, with a limit of quantification of 10 ppb. Cowell and Steinmetz (1990) found that, of 96 urine samples analyzed from silvicultural workers, only 5 contained quantifiable levels of glyphosate. The highest measurement was 14 ppb. In a recent pilot study with three farmers and their families, there were no quantifiable residues of glyphosate in the study except one farmer with a urinary glyphosate measurement of 12 ppb on the day of a 5-hour, hand-wand sprayer application to weeds along a fence line (Alexander et al., 1999). This application method is similar to those in the previous silvicultural studies. In a worst case analysis, Williams et al. (2000) estimated that an adult worker’s peak acute exposure to glyphosate during application was 56.2 g(kg body weight) day and, for a 5-day working week, the chronic applicator exposure was 8.5 g(kg body weight)day. Comparison of these values to lowest relevant NOEL of 175 mg/kg/ day in a the rabbit developmental toxicity study produced margins of exposure (MOEs) of 3114 and 20,588 in acute and chronic exposure, respectively. Actual exposures are anticipated to be significantly less. Jauhiainen et al. (1991), in addition to biological and inhalation monitoring of glyphosate of forestry workers during application, had

Chapter | 92  Inhibitors of Aromatic Acid Biosynthesis

each worker receive a medical examination on the first and last days that Roundup® herbicide was applied and a follow-up examination 3 weeks after the last application day. No changes were noted in hematology, clinical chemistry, electrocardiogram, pulmonary function, blood pressure, or heart rate. Accidental exposures to small volumes of glyphosate have not produced serious effects. In spite of this experience, it has been stated that glyphosate is a leading cause of pesticide poisoning in California. California’s Department of Pesticide Regulation (CDPR) pesticide incident program accepts telephone inquiries from physicians, who are required to report pesticide incidents, as well as from the general public. Although many calls are purely informational or report effects limited to topical irritation, all telephone calls are recorded as “poisonings,” incorrectly suggesting some degree of systemic illness. In 1994, CDPR reported only 13 “definite” or “probable” calls related to glyphosate exposure alone among the 1995 total calls received (California Environmental Protection Agency, 1996). Eleven of these 13 cases reported only minor and reversible eye irritation likely due to accidental overexposure. Of the remaining two cases, one involved a worker who reported a headache in addition to eye irritation. The other case involved symptoms related to ingestion and/or aspiration of hydrocarbon solvent. The latter case cannot be related to glyphosate itself or to a marketed formulation, because commercial preparations of glyphosate are not formulated using hydrocarbon solvents. The CDPR noted in its 1994 Pesticide Illness Surveillance Report that greater than 80% of the people affected by glyphosate experienced only irritant effects and that, of the 515 pesticide-related hospitalizations recorded over the 13 years on file, none was attributed to glyphosate. Statements to the effect that glyphosate is a major cause of clinical poisoning in California are clearly not substantiated by the available data. It has become customary to generically refer to any organic compound containing phosphorous as an “organophosphate.” However, there are actually different classes of “organic phosphate” compounds that are determined by the atoms attached to the phosphorus. The phosphorus atom of a true organophosphate is attached only to oxygen atoms (O’Brien, 1967). The structure of glyphosate is different in two important respects. First, the phosphorus atom is attached to the remainder of the molecule by a carbon atom, not an oxygen. This classifies the glyphosate molecule as an organophosphonate (O’Brien, 1967). Second, there are no other side chains attached to the phosphorus atom. Glyphosate consists of a glycine moiety and a phosphonomethyl moiety. These distinctions are important for the following reason. The nature of the groups attached to the phosphorous determine how strongly the molecule will interact with the enzyme cholinesterase (O’Brien, 1967). The groups attached to the phosphorus of a true organophosphate allow it to be readily hydrolyzed by cholinesterase.

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In contrast, the carbon-phosphorus bond of an organophosphonate is not easily broken (Roberts and Caserio, 1965). This interaction is responsible for the subsequent inhibition of the enzyme and disruption of normal nerve function. The phosphorus of an organophosphonate such as glyphosate, on the other hand, does not react with or inhibit cholinesterase. Thus, this type of molecule does not interfere with normal nerve function (Williams et al., 2000). Large amounts of glyphosate-based herbicides are occasionally deliberately ingested to attempt suicide and may result in serious gastrointestinal, cardiovascular, pulmonary, and renal effects and possibly death. Aggressive supportive care is recommended (Tominack et al., 1989). Glyphosate is sometimes mistakenly referred to as an organophosphate, thus contributing to the incorrect perception that glyphosate is a cholinesterase inhibitor similar to the organophosphate insecticides. As a result, some patients who have ingested glyphosate-based herbicides have received inappropriate medical treatment which may have worsened their condition. Atropine or 2-PAM (Pralidoxime) are not indicated in the treatment of glyphosate exposure.

References Acquavella, J. F., Weber, J. A., Cullen, M. R., Cruz, O. A., Martens, M. A., Holden, L. R., Riordan, S., Thompson, M., and Farmer, D. R. (1999). Human ocular effects from self-reported exposure to Roundup herbicides. Hum. Expl. Toxicol. 18, 479–486. Alexander, B. H., Mandel, J. S., Baker, B. A., and Honeycutt, R. (1999). “The Farm Family Exposure Pilot Study.” Unpublished draft final report. Bodden, R. M. (1988). “Metabolism Study of Synthetic 13C/14C-Labeled Glyphosate and Aminomethylphosphonic Acid in Laying Hens.” Unpublished report, Hazleton Laboratories America, Inc., Madison, WI. Branch, D. K. (1981). “Primary Eye Irritation of Isopropylamine Salt of Glyphosate to Rabbits Eyes.” Unpublished report, Monsanto Environmental Health Laboratory, St. Louis, MO. Busch, B. (1987). “Primary Eye Irritation Study of Monosodium Salt of Glyphosate in New Zealand White Rabbits.” Unpublished report. Food and Drug Research Laboratories, Inc., Waverly, NY. California Environmental Protection Agency (1996). “California Pesticide Illness Surveillance Program Summary Report 1994.” Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento. Colvin, L. B., and Miller, J. A. (1973). “Residue and Metabolism—The Gross Distribution of N-Phosphonylmethylglycine-14C in Rabbits.” Unpublished report, Monsanto Company, St. Louis, MO. Cowell, J. E., and Steinmetz, J. R. (1990). “Assessment of Forest Worker Exposures to Glyphosate During Backpack Foliar Applications of Roundup® Herbicide.” Unpublished report, Monsanto Company, St. Louis, MO. Dudek, B. R. (1987). “Acute Toxicity of Rodeo® Herbicide Administered by Inhalation to Male and Female Sprague-Dawley Rats.” Unpublished report. Monsanto Environmental Health Laboratory, St. Louis, MO.

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Franz, J. E., Mao, M. K., and Sikorski, J. A. (1997). “Glyphosate: A Unique Global Herbicide,” ACS Monograph No. 189. Am. Chem. Soc., Washington, D.C. Jauhiainen, A., Rasanen, K., Sarantila, R., Nuntineg, J., and Kangas, J. (1991). Am. Ind. Hyg. Assoc. J. 52, 61–64. Lavy, T. L., Cowell, J. E., Steinmetz, J. R., and Massey, J. H. (1992). Conifer seedling nursery worker exposure to glyphosate. Arch. Environ. Contam. Toxicol. 22, 6–13. Maibach, H. I. (1986). Irritation, sensitization, photoirritation, and photosensitization assays with a glyphosate herbicide. Contact Dermatitis 15, 152–156. National Toxicology Program (1992). “Technical Report on Toxicity Studies of Glyphosate (CAS No. 1071-83-6) Administered in Dosed Feed to F344/N Rats and B6C3F, Mice.” Toxicity Report Series Number 16, NIH Publication 92-3135, July 1992, National Toxicology (NTP), U.S. Department of Health and Human Services, Research Triangle Park, NC. O’Brien, R. D. (1967). Organophosphates: Chemistry and inhibitory activity. In “Insecticides: Action and Metabolism,” pp. 32–54. Academic Press, New York. Ridley, W.P., and Mirly, K. (1988). “The Metabolism of Glyphosate in Sprague-Dawley Rats. I. Excretion and Tissue Distribution of Glyphosate and Its Metabolites Following Intravenous and Oral Administration.” Unpublished report. Monsanto Environmental Health Laboratory, St. Louis, MO.

Hayes’ Handbook of Pesticide Toxicology

Roberts, J. D., and Caserio, M. S. (1965). Organophosphorus compounds. In “Basic Principles of Organic Chemistry,” pp. 1194–1215. Benjamin, New York. Tominack, R. L., Conner, P., and Yamashita, M. (1989). Clinical management of Roundup herbicide exposure. Jpn. J. Toxicol. 2, 187–192. U.S. Environmental Protection Agency (U.S. EPA) (1992). Pesticide tolerance proposed rule. Fed. Reg. 57, 8739–8740. U.S. Environmental Protection Agency (U.S. EPA) (1993). “Re­registration Eligibility Decision (RED): Glyphosate.” Office of Prevention, Pesticides and Toxic Substances, U.S. Environmental Protection Agency, Washington, D.C. U.S. Environmental Protection Agency (U.S. EPA) (1999). Glyphosate: Pesticide tolerance, Final Rule—40 CFR, Part 180 [Opp-300835; FRL-6073-5]. Fed. Reg. 64(71), 18360–18367. Wester, R. C., Melendres, J., Sarason, R., McMaster, J., and Maibach, H. I. (1991). Glyphosate skin binding, absorption, residual tissue distribution, and skin decontamination. Fundam. Appl. Toxicol. 16, 725–732. World Health Organization (WHO) (1994). “Glyphosate,” Environmental Health Criteria No. 159. International Programme of Chemical Safety (IPCS), World Health Organization, Geneva. Williams, G. M., Kroes, R., and Munro, I. C. (2000). Safety evaluation and risk assessment of the herbicide roundup and its active ingredient, glyphosate, for humans. Regul. Toxicol. Pharmacol. 31, 117–165.

Section XIV

Other Selected Pesticides

(c) 2011 Elsevier Inc. All Rights Reserved.

Chapter 93

Toxicology of DDT and Some Analogues Andrew G. Smith Medical Research Council Toxicology Unit, Leicester, United Kingdom

93.1  Introduction Chlorinated insecticides continue to be regarded as pernicious environmental pollutants. Even the few uses of these insecticides licensed worldwide and the few countries this applies to are under tremendous pressure to cease completely. In contrast, it is difficult for us now to understand that in their time the chlorinated insecticides were of outstanding importance for human health. Like penicillin, wartime showed their tremendous utility. With their introduction, pests could be confidently controlled and importantly, eradicated. It is estimated that DDT alone was responsible for the saving of tens of millions of human lives. Many areas of the world are now free of such diseases as malaria purely as the result of DDT used on lice and mosquitoes in the period 1942–1960s. Our present concerns of environmental toxicity of these chemicals derive from the lack of appreciation that the large-scale administration of these persistent chemicals might have significant costs for wildlife. Whether there are any costs for chronic human health through environmental and dietary exposure is a very difficult issue. In the last 10 years or so, DDT, its analogues, and other chemicals have been considered as possible endocrine disrupters causing cancer and reproductive disorders. However, on the whole, DDT-type chlorinated insecticides have apparently been extremely safe for humans. With unknown repercussions of climate change, it would seem prudent to retain the use of DDT, albeit in a tightly controlled manner, as an important reserve weapon in the fight against insect pests that are vectors for human diseases and might proliferate in changing conditions (Mandavilli, 2006a,b). Resistance to newer chemicals may also develop. It is not always easy to substitute with newer insecticides that are more expensive and have their own undesirable effects. This chapter describes Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

aspects of the known toxicity of DDT and its analogues such as methoxychlor. Although much of the work reported is now decades old, it is extremely important and very pertinent to today’s discussions and arguments on the hazards of environmental exposures (Smith, 2000).

93.2  DDT DDT came to widespread attention because it dramatically controlled typhus and malaria in wartime. When it became available for civilian use, it was used to control flies and other pests that annoy large numbers of people and may transmit disease, and it increased the production of important crops. Knowledge that traces of it are stored in essentially everyone in the world has kept DDT in the spotlight. Later it was implicated in the injury of a wide variety of wildlife. Under these circumstances, it is no wonder that DDT probably has been studied more thoroughly than any other pesticide and is used to illustrate many principles and concepts in toxicology including very important topics such as human exposure levels and effects on domestic and wild animals. In the following discussion, details of storage and excretion of DDT in humans are covered as well as toxic actions in humans and experimental animals. The topic is now so huge that only outlines of metabolism, tissue levels and in vitro effects can be reported here. DDT was first synthesized by Zeidler in 1874; however, it was put to no use until its insecticidal properties were demonstrated by Paul Müller in 1939. The first sample sent to the United States arrived in September 1942. Results of tests in the UK and United States were so encouraging that manufacture was given high priority. At first, the entire production was used for the protection of troops against malaria, typhus, or certain other vector-borne diseases, or 1975

Hayes’ Handbook of Pesticide Toxicology

1976

against biting flies or other insects that are merely pests (Hayes, 1982). As the supply increased, DDT was used in the United States for control of malaria in the vicinity of military installations, ports, and transportation centers. As a result of this effort, mosquito transmission of malaria was brought to an end in the United States in 1953, even though military personnel and other infected persons from the tropics continued to reintroduce the disease extensively as late as 1972 and in diminishing numbers thereafter. The revolution in the control of malaria and typhus among allied troops and among certain civilian populations during World War II was accomplished with relatively little DDT. Far greater amounts were required for the control of agricultural and forest pests that became possible after the compound was released in the United States for commercial use on August 31, 1945. Civilian use in other countries became possible a little later with tremendous effect. An account of its discovery and early use, especially in Europe, can be found in West and Campbell (1946).

United Kingdom as dicophane (BP), in Sweden as klorfenoton, and in the United States as chlorphenothane (USP). DDT was sold under a variety of trade names, including Anofex, Cesarex, Didimac, Digmar, Diniocide, Genitox, Guesarol, Gyron, Ixodex, Neocid, and Zerdane. Code designations for DDT include IMS-16 and ENT-1,506. The CAS registry number for p,p-DDT is 50-29-3. DDT has the empirical formula C4H9Cl5 and a molecular weight of 354.49. Pure p,p-DDT is a white, tasteless, almost odorless crystal-line solid melting at 108.5 to 109.0°C. Technical DDT is a waxy solid. A typical example of technical DDT had the following composition: p,p-DDT, 77.1%; o,p-DDT, 14.9%; p,p-DDD, 0.3%; o,p-DDT, 14.9%; p,p-DDD, 0.3%; o,p-DDD, 0.1%; p,p-DDE, 4.0%; o,pDDE, 0.1%; and unidentified compounds, 3.5%. The vapor pressure of DDT is 1.5  107 mmHg at 20°C. DDT is highly soluble in a polar organic solvents: solubility per 100 ml acetone, 58 g; ethanol, 2 g; benzene, 106 g; carbon tetrachloride, 45 g; cyclohexanone, 116 g; ethyl ether, 28 g; petroleum ethers and kerosene, 4–10 g. It is practically insoluble in water. The structure of p,p-DDT and the structures of some of its analogues are compared in Figure 93.1. A more complete list can be found in previous editions of this chapter (Smith, 1991, 2001). It must be emphasized that even the commercially available insecticidal analogues have strikingly different properties. Especially remarkable are the slow metabolism and marked storage of DDT and its metabolite DDE and the rapid metabolism and negligible storage of methoxychlor. Figure 93.1 does not include the wide range of compounds that have been synthesized and studied in connection with structure–activity relationships, often with the hope of emphasizing the good properties of DDT and reducing its undesirable properties. For a more extensive consideration of analogues see Metcalf (1973).

93.2.1  Identity, Properties, and Uses p,p-DDT is 1,1-(2,2,2-trichloroethylidene)-bis(4-chlorobenzene). Common nomenclatures that have been used are 1,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane, 1,1,1-trichloro2,2-bis(4-chlorophenyl)ethane, and 1,1-bis(4-chlorophenyl)2,2,2-trichloroethane. Because the older terminology has been used widely in the past and continues to be so, especially since many abbreviations are based on it (e.g., p,p-DDT and o,p-DDT), the o and p nomenclature will be used for referring to DDT in its abbreviated form. DDT is universally accepted as the common name of the insecticide. As approved by BSI, DDT refers to the technical product, and there is historical justification for that practice because DDT is an acronym for dichlorodiphenyltrichloro­ ethane. p,p-DDT was approved by BSI as a separate term. Zeidler called the compound di-monochlorophenyltrichloräthan. When used as a drug, DDT was known in the CCl3

Cl

1

93.2.2  Formulations and Production Technical DDT has been formulated in almost every conceivable form including solutions in xylene or petroleum CHCl2

CHCl2

Cl

Cl

Cl 2

CCl3

CH3CH2

COOCH2CH3

CCl3

OH CH3O

4

OCH3

Cl

5

CH2CH3

3

OH Cl

Cl

6

Cl

Figure 93.1  Structures of DDT and some of its derivatives that have been used or are still in use. 1. p,p-DDT. 2. TDE. 3. Ethylan. 4. Methoxychlor. 5. Chlorobenzilate. 6. Dicofol. The names used here are those which are commonly encountered. TDE as a metabolite of DDT is usually called DDD. As a drug, the o,p-isomer is called Mitotane. Further analogues can be found in previous editions of this chapter (Smith, 1991, 2001).

Chapter | 93  Toxicology of DDT and Some Analogues

distillates, emulsifiable concentrates, water-wettable powders, granules, aerosols, smoke candles, charges for vaporizers, and lotions. Aerosols and other household formulations were combined with synergized pyrethrins. Production and use of DDT in the United States and other countries have been discussed previously (Hayes, 1991; Smith, 1991). Before 1945, all of the DDT produced in the United States was used or allocated by the military services for medical and public health uses. Early in 1945, it became available for extensive experimental work in agriculture, and it was commercially available in limited quantities early in the autumn of the same year. The results were so spectacular that use increased until 1959. In response to a demand for exports, production continued to increase until about 1963. Even before 1963, some restrictions were placed on its use, mainly to minimize residues in food and in the feed of animals that produced milk and meat. Among the first of these restrictions was that on its use in the dairy cattle industry. Another important factor reducing the use of DDT was the increasing resistance of pests. One of the first species to be affected was the housefly; because of its abundance and widespread distribution, its resistance was bound to be noticed by the public generally. Although many pests of public importance have been resistant to DDT in some or all of their range, resistance among vectors of malaria has been minimal. Because malaria control constitutes such a large segment of vector control, the use of DDT for vector control remained stable for many years, while its use in agriculture continued to decline, especially in temperate climates. Prophetically when Sweden banned DDT from January 1, 1970, they pointed out that “the need for insecticides is rather small in Sweden compared to that in many other countries” and that the ban of this and certain other chlorinated hydrocarbon insecticides could be used as a tool to explore scientific problems about their movement (Hayes, 1982). In order to respond to ecologists who considered that the widespread occurrence of DDT in the environment was inherently bad and was the direct cause of injury to certain fish and birds, government agencies of some other countries attempted to justify restrictions on the use of DDT by its alleged threat to human health. This did not prevent the same agencies from providing that DDT might be used, if needed, to combat any future threat from vector-borne disease within their boundaries. To this day the hazard of DDT and the persistent metabolite DDE to humans is still a highly debatable matter. Although many countries severely restrict or ban the use of DDT, it is still used for both agriculture and vector control in some tropical countries. It is possible that complete abolishment of its use worldwide in vector control might have significant repercussions. Information apparently is not available on how much agricultural use involving food protection still occurs. How much of the little use of DDT is in public health is still also unknown, but the picture with malaria control is clear. In

1977

1971, WHO calculated that substitution of malathion or propoxur for DDT would increase the cost of malaria control approximately 3.4- and 8.5-fold, respectively, and this increase could not be supported in some countries without a decrease in the coverage of control programs (WHO, 1971). Despite these increased costs, DDT use has been mainly overtaken by other pesticides. However, a proposal for its complete ban was a controversial matter (Attaran et al., 2000; Curtis and Lines, 1999; Roberts et al., 2000). In 2001, the United Nations Development Programme agreed that use of DDT could continue in circumtances to protect against malaria, especially in bed nets, where resistance to other pesticides had arisen.

93.2.3  Toxicity to Laboratory Animals 93.2.3.1  Symptomatology The description of DDT intoxication in animals given by Domenjoz remains one of the best (Domenjoz, 1944). The first perceptible effect is abnormal susceptibility to fear, with violent reaction to normally subthreshold stimuli. There is definite motor unrest and increased frequency of spontaneous movements. As poisoning increases, hyperirritability like that seen in strychnine poisoning develops, but convulsions do not appear at this time. A fine tremor, recognizable at first only as a terror reaction, is later present as an intention tremor in connection with voluntary movement, and then intermittently without observable cause. Finally, it is present as a coarse tremor without interruption even for several days. Spontaneous movement is limited, and food intake stops so that surviving animals lose weight. In the later stages, especially in some species, there are attacks of epileptiform, tonic-clonic convulsions with opisthotonos. All the signs are strengthened by external stimuli and become manifest at first through external stimuli. In all stages, the animals show normal position and labyrinth reflexes. The picture of poisoning in mammals recalls the disturbances of movement and tone that are known in human pathology as the amyostatic syndrome. Symptoms appear several hours after oral administration of the compound, and death may follow after 24–72 h. The latent period after intravenous administration at about the LD50 level is approximately 5 min; signs of poisoning reach a maximal level in about 30 min, and survivors are symptomfree in 18–24 h. Animals that survive recover completely. In addition to the features of poisoning already mentioned, Cameron and Burgess (1945) noticed that as rats, guinea pigs, and rabbits become sick they become cold to the touch and show ruffled fur. Some show diarrhea. Muscular tremors were preceded by muscular weakness, which occurred first in the back and later in the hind legs. The front legs were relatively spared so that animals showing marked weakness of the hindquarters could still drag themselves about. However, several authors have found

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1978

that the tremor characteristic of DDT poisoning generally starts in the muscles of the face, including the eyelids, and spreads caudally with variable severity until all the muscles are affected. Furthermore, although weakness of hindquarters has been seen, it was not a common finding. Although there is a general similarity in the clinical effects of DDT in all vertebrate species, some characteristic differences exist. Cats show greater extensor rigidity and opisthotonos than other laboratory animals. The stiffness appears first in the distal part of the extremities and later extends to the proximal part and to the trunk. Poisoned cats show marked pilomotor activity. Convulsions are also prominent in dogs, as is ataxia. Tremors are so pronounced in rats that it may be difficult to detect clonic convulsions in them. Poisoning produced by repeated doses of DDT differs from that produced by a single dose only insofar as the animal may be gradually debilitated, especially by malnutrition. If food intake is maintained, tremor may last for weeks or even, intermittently, for months. If the animals survive a short time after dosing stops, recovery is complete. However, food intake may be interfered with in at least two ways. Tremor and more severe signs may interfere mechanically with eating. Animals offered food containing high concentrations of DDT often eat little or

nothing and lose weight rapidly. This seems to be due to taste, not an effect on appetite, as the same animals will show excellent appetites when offered the same kind of food containing no DDT just after refusing the major portion of their daily ration of contaminated food. Animals that have suffered severe weight loss as a result of DDT poisoning may die partly as a result of general debility. Even though severely ill, animals that survive a few days after the last of many doses of DDT go on to recovery. Table 93.1 summarizes the acute toxicity of DDT to common laboratory animals. It may be concluded that dissolved DDT is absorbed by all routes, although DDT powder is absorbed through the skin to only a negligible degree. Remarkably, it is frequently impossible to put enough DDT dust on the skin of animals to kill them, so that an LD50 value for this formulation cannot be determined with precision by the dermal route. Although formulation is important in determining the toxicity of DDT by other routes, the difference is not so great as it is in connection with skin exposure. In round figures, DDT is about four times more toxic when given intravenously than when given orally and about 40 times more toxic intravenously than dermally. In these studies, DDT, like some other lipophilic chemicals, was more toxic orally as a solution in vegetable oil or animal fat

Table 93.1  Comparison of Acute LD50 to Laboratory Animalsa Species

Formulationb

S.c. (mg/kg)

Rat

w/p

2000

o

200–1500

w/p o

Mouse

Guinea pig

Dermal (mg/kg)

500–2500

1000

113–450

250–3000

1000–1500

300–1600

375

300

100–800

250–500

2000

1500

250–560

1000

275

375

300–1770

300–2820

900

250–3200

68

650

32

w/p

Modified from Hayes (1959). w/p suspension in water or as powder, o, solution in oil.

b

2100

300 100–410

o a

30–41

w/p o

Monkey

150

w/p o

Cat

80–200

w/p o

Dog

Oral (mg/kg)

47

I.p. (mg/kg)

w/p o

Rabbit

I.v. (mg/kg)

55

Chapter | 93  Toxicology of DDT and Some Analogues

than when it was given in some petroleum fractions. Acute oral LD50 values of DDT metabolites commonly found in tissues or excreta are less toxic than the most absorbable preparations of the parent compounds. At an oral dosage of 150 mg/kg, p,p-DDT produces severe illness in rats and kills about half of them, but o,pDDT at the same dosage produces no illness, even though the concentrations of the two compounds in the brain at various intervals after dosing are about the same. At a dosage of 3000 mg/kg, o,p-DDT produces mild to moderate illness, and the concentration in the brain is five to nine times the concentration of p,p-DDT necessary to produce similar symptoms. Thus, p,p-DDT appears to be inherently more toxic than the o,p isomer (Dale et al., 1966). Rats tolerate higher tissue levels of DDA than of DDT. Eighteen hours after intravenous injection of DDA at a rate of 100 mg/kg, tissue levels still were higher than are usually found in animals fatally poisoned by DDT (Judah, 1949). DDA produces somewhat less injury than DDT to the liver but, especially at high intravenous dosages, produces greater damage to the kidney (Lillie et al., 1947). This is consistent with the finding of Spicer et al. (Spicer et al., 1947) that, following administration of DDT, DDA constitutes a higher proportion of DDT-related compounds in the kidney (25%) than in any other tissue, being 12% in the liver, 10% in the brain, and even less in other tissues. Young animals eat more than adults in relationship to their body weight. For this and other reasons, young animals often are more susceptible than adults to chemicals

Table 93.2  Effect of Age on the Toxicity of DDT to Rats Number of doses

Agea

LD50 (mg/kg)b

1

Newborn

4000

1

Newborn

2356

1

10 days

728

1

14–16 days

437

1

Weaning

355

1

2 months

250

1

2 months

152

1

3–4 months

194

1

Middle-aged

235

1

Adult

225

4

Preweaning

279

4

Adult

285

Original data can be found in Harrison, 1975; Henderson and Woolley, 1969; Lu et al., 1965; Mitjavila et al., 1981. a Data from more than one strain of rat. b Total intake of one or more doses.

1979

in food. However, there is no evidence that DDT is more toxic to young animals of any species, including humans, and in the rat the young are less susceptible to a single dose (Table 93.2). They are about equally susceptible to repeated doses. According to Henderson and Woolley (Henderson and Woolley, 1969), the relative insusceptibility of the young is associated with relatively poor absorption of DDT by their central nervous systems and by lesser inherent susceptibility of the young brain to DDT already absorbed by it. Further study showed that fatal poisoning of both 10- and 60-day-old rats involves hyperexcitability and intense tremor followed by prostration and eventual respiratory failure (Henderson and Woolley, 1970). However, in the adult rat, DDT causes convulsions, an increase in respiration and heart rate, and a lethal increase in body temperature prior to death, but the body temperature of the immature rat decreases during acute intoxication by DDT. The authors suggested that, whereas DDT is a direct depressant of respiration in both young and old rats, the additional toxic responses manifested by seizures and hyperthermia account for the increased lethality of DDT in mature animals. No acute LD50 could be established for hamsters (Agthe et al., 1970), which also seem resistant to chronic effects of DDT (Table 93.3). There is virtually no sex difference in the acute toxicity of DDT to rats; the reported LD50 is 113 and 118 in males and females, respectively (Gaines, 1960). A similar situation was observed with mice (Agthe et al., 1970). When DDT is fed to rats at ordinary dietary levels, the two sexes store it equally. However, at higher dosages, females store more of the compound; the difference is probably explained mainly by the lower activity of the hepatic cytochrome P450 enzymes affecting the rate of metabolism in female rats and only in part by relatively higher food intake of the females.

93.2.3.2  Response to Repeated Doses The effects of repeated doses of DDT are summarized in Table 93.3. The 90-dose oral LD50 of technical DDT in rats is 46.0 mg/kg/day (Gaines, 1969). The chronicity index is 5.4. Thus the compound has only a moderate tendency to cause cumulative effects, and this limited tendency is fully explained by the accumulation of DDT itself in tissues as a result of containing intake. In fact, this accumulation, which is strictly dosage-dependent, is detectable at all measurable levels of intake. If storage is considered undesirable per se, then DDT is without a no-injurious-effect level. However, the same may be said for all compounds that are absorbed, for the presence of all of them in the bodies of exposed organisms – perhaps at very low levels and for relatively short periods – may be assumed; failure to demonstrate low levels of storage may not depend on physiology but only on limitations of the analytical techniques employed. Fat storage can be viewed as a protective mechanism.

Hayes’ Handbook of Pesticide Toxicology

1980

Table 93.3  Effect on Various Species of Prolonged Oral Administration of DDT Range Dosage or (mg/kg/day) concn in diet (ppm)

Species,a number, and sex

Maximum duration

Results

References

41–80

800 ppm

Rat

2 years

Increased mortality, typical liver changes, and liver carcinomas

A

46 mg/kg, then 140 ppm

Mouse 36 M, 36 F

1.5 yrs

Hepatomas in 51 and 21% of M and F compared with 18 and 0.6% of controls

B

1000 ppm

Hamster 25 M, 30 F

1.9 years

No liver tumors and survival slightly less than controls

C

1000 ppm

Hamster 30 M, 30 F

1.5 years

No liver tumors but decreased serum cholinesterase

D

1000 ppm

Hamster 35 M, 36 F

2.4 years

No liver tumors and survival as controls

E

3200 ppm

Dog 10

4 years

100% Mortality; liver damage, no tumors

F

5000 ppm

Monkey 1 M

10 weeks

100% Mortality

G

50 mg/kg/day

Monkey 6

14 weeks

100% Mortality; no hematologic effects

H

400 ppm

Rat 24 M, 12 F

2 years

Increased mortality, typical liver changes

A

500 ppm

Rat 37 M, 35 F

2.9 years

Liver tumors in 45%

I

500 ppm

Rat 38 M, 38 F

2.3 years

Liver tumors in 18% F

J

250 ppm

Mouse 103 M, 90 F

2 gen

Risk of liver tumor increased 3.7- and 18.5-fold in M and F, respectively

K

250 ppm

Mouse 31 M, 121 F

2 gen

Liver tumors in 48 and 59% of M and F

L

500 ppm

Hamster 39 M, 40 F

1.7 years

No liver tumors and survival as controls

M

2000 ppm

Dog

4 years

25% Mortality; minor liver damage but no tumors

F

100 ppm

Mouse 100 M, 100 F

2 years

Hepatomas increased in F of one strain but no increase in hepatocarcinomas

N

100 ppm

Mouse 30 M, 30 F

2 years

Risk of liver tumors increased 4.4-fold

O

100 ppm

Mouse 30 M, 3 F

2 years

Risk of liver tumors increased 3.3- and 4.2-fold in M and F, respectively

P

50 ppm

Mouse 127 M, 104 F

2 gen

Risk of liver tumors increased 2.45- and 3.46-fold in M and F, respectively

K

50 ppm

Mouse 30 M, 30 F

2 years

Risk of liver tumors increased 2.9-fold

O

400 ppm

Dog 2

4 years

No effect

F

20 ppm

Mouse 48 M, 128 F

2 gen

No increase in tumors

L

200 ppm

Monkey

7.5 years

No effects

G

1.26–2.5

10 ppm

Mouse 104 M, 124 F

2 gen

Risk of liver tumors increased 2.26- and 2.46-fold in M and F, respectively

K

0.63–1.26

25 ppm

Rat

2 years

No clinical effect; M survived longer than controls

Q

0.31–0.63

10 ppm

Rat

2 years

Typical liver changes; no effect on reproduction

R

12.5 ppm

Rat

2 years

No effect

Q

2.8–3.0 ppm

Mouse 683

5 gen

Tumors in 28.7%, including lung carcinomas, lymphomas, and leukemias

S

21–40

11–20

6–10

2.6–5

Chapter | 93  Toxicology of DDT and Some Analogues

1981

Table 93.3  (Continued) Range Dosage or (mg/kg/day) concn in diet (ppm)

Species,a number, and sex

0.16–0.31

2 ppm

Mouse 124 M, 111 F

2 gen

Risk of liver tumor doubled in M, unchanged in F

K

2 ppm

Mouse 58 M, 135 F

2 gen

No increase in tumors

L

2.5 ppm

Rat

2 years

No effect

Q

0.08–0.16

Maximum duration

Results

References

References: A, Fitzhugh and Nelson, 1947; B, Innes et al., 1969; C, Agthe et al., 1970; D, Graillot et al., 1975; E, Rossi et al., 1983; F, Hayes, 1982, Lehman, 1965; G, Durham et al., 1963, 1963; H, Cranmer et al., 1972; I, Rossi et al., 1977; J, Cabral et al., 1982a; K, Tomatis et al., 1972; L, Terracini et al., 1973; M, Cabral et al., 1982b; N, Fitzhugh, 1970; O, Walker et al., 1973; P, Thorpe and Walker, 1973; Q, Treon and Cleveland, 1955; R, Fitzhugh, 1948; S, Tarjan and Kemeny, 1969. a Various strains of rats were used and can be found in the individual papers or Hayes (1982); Smith (1991). b Slides reexamined by Reuber, 1978.

A number of papers have reported no-effect levels for DDT within parameters other than storage, namely rat, 0.05 mg/kg/day; dog, 8 mg/kg/day; and monkey, 2.2–5.54 mg/kg/day (see Smith, 1991). There remain reports of effects in animals at the lowest dosages investigated. For example, decreased serum albumin and increased - and -globulins in the blood of rats and rabbits maintained on a dosage of 0.2 mg/kg/day for 3–11 months were reported by Kagan et al. (1969). In summary, the lowest dosages that have been studied in animals are of the same order of magnitude as those encountered by people who made or formulated DDT and, therefore, hundreds of times greater than the dosages encountered by ordinary people. The animal studies have continued long after a steady state of storage has been achieved. The results permit the conclusion that bioaccumulation sufficient to produce neurotoxicity or other clear clinical effects, including a reduction of the life span, can occur only at dosage levels substantially higher than those encountered by the most heavily exposed workers, let alone those exposed environmentally. DDT dosages encountered by workers produce in some groups of mice and rats a small but detectable increase of the liver changes (hypertrophy, margination, and lipospheres) characteristic of rodents at much higher doses (Smith, 1991). The same changes occur in low incidence in control mice and rats but not in other animals.

93.2.3.3  Absorption Most DDT dust is of such large particle size that any that is inhaled is deposited in the upper respiratory tract and eventually is swallowed. Toxicity data indicate that respiratory exposure to DDT is of no special importance. The absorption of DDT from the gastrointestinal tract is slow. Whereas intravenous injection at the rate of 50 mg/kg produces convulsions in rats in 20 min, convulsions occur only after 2 h when DDT is administered orally at a rate two or more times the LD50 value. The onset of convulsions

is delayed for about 6 h when DDT is given to rats orally at approximately the LD50 value (Dale et al., 1963). DDT dissolved in animal or vegetable fats is absorbed from the gastrointestinal tract about 1.5–10 times more effectively than is undissolved DDT, (for example, Keller and Yeary, 1980; Palin et al., 1982), but large doses of the compound in the gastrointestinal tract are poorly absorbed from nonabsorbable solvents. At high dosage levels, less [14C]DDT is absorbed and stored in organs following oral than following intraperitoneal administration, and a higher proportion is excreted in the feces than after intraperitoneal administration (40 vs. 0.9%) (Bishara et al., 1972). However, in connection with small repeated doses, the kind of solvent used made little difference; probably the occurrence of bile in the intestine and the presence of some fat in the diet are sufficient to promote absorption of the compound. Rothe et al. reported that after giving radioactive DDT by stomach tube as an emulsion of a peanut oil solution they recovered 41–57% of it in lymph. Less than 0.1% was found in the urine, 7.4–37.1% was in the feces or in the intestinal contents, and 19–67% of the activity was found in the carcass (Rothe et al., 1957). Of the administered DDT not found in feces and intestinal contents, 47–65% was found in the lymph. Fifty percent of the DDT-derived material found in the lymph was absorbed in the first 2.5–7 h, and 95% was absorbed by 18 h. Because the lymphatic duct in the rat is not a single vessel, Rothe et al. were unable to exclude the possibility that some or all of the DDT that they later recovered from the carcasses of their animals had been transported to the general circulation by collateral lymph vessels rather than by the hepatoportal system. They gave indirect evidence for supposing that little or no DDT is absorbed from the gastrointestinal system by the blood, and this has been confirmed (Palin et al., 1982). Most of the DDT absorbed into the lymph is carried in the lipid core of chylomicrons and thence into the plasma proteins (Pocock and Vost, 1974; Sieber et al., 1974). DDT seems to be transferred to the tissues from chylomicrons at a faster rate, and in a different distribution,

1982

to cholesteryl esters (Trevaskis et al., 2006a,b). p,p-DDT is taken up at a rate which is different from those of its metabolites and o,p-DDT (Sieber, 1976) and which does not strictly parallel differences in lipid solubility. As already stated, dermal absorption of DDT is very limited.

93.2.3.4  Distribution and Storage The distribution and storage of DDT in animals can be summarized as below. Original references can be found in Smith (1991). 1. DDT is stored in all tissues, the highest concentrations of DDT usually being found in adipose tissue. Rats store DDT in their fat at all measurable dietary levels including trace concentrations. 2. Following repeated doses, storage in the fat increases rapidly at first and then more gradually until a peak or plateau is reached. Repeated doses at a moderate rate could result in greater total storage of DDT in the fat than a single dose at the highest rate that can be tolerated or even a single dose at a rate that frequently is fatal. The equilibrium storage of DDT in each tissue varies directly with the daily dosage. However (with the apparent exception of the dog), storage in the fat and perhaps in other tissues is less extensive in relation to dosage at higher dietary levels. 3. The rat apparently tends to lose a part of the DDT it has stored in fat at the peak level reached in about 6 months, even though it continued on the same diet. 4. There is a measurable difference between the storage patterns of different species; that of the dog differs most. 5. At higher dosage levels but not at ordinary residue levels, the female rat consistently stores more DDT in its fat than the male when fed the same diet. The difference is accounted for only in part by the greater food intake of the female and must depend partly on more rapid biotransformation in the male. Other species show little or no sex difference. 6. The amount of DDT stored in the tissues is gradually reduced if exposure to the compound is discontinued or diminished. Other observations regarding storage include the finding that rats whose brains contain DDT at a concentration of 25 ppm or less (wet weight) usually survive, whereas higher levels tend to be fatal regardless of whether absorption followed one or many doses. Of samples that may be collected in vivo, the concentration of DDT in serum most accurately reflects its concentration in the brain, the critical tissue. Adams et al. observed that about the same concentrations of DDT and related compounds are stored by male rats and by females that reproduce successfully (Adams et al., 1974). The lower storage in mated females probably

Hayes’ Handbook of Pesticide Toxicology

is accounted for by transfer to the young via the placenta and the milk. This is now considered an important factor in the risk assessment of other lipophilic chemicals such as methyl mercury and dioxins. When DDT, some of its analogues, and several other chlorinated hydrocarbon insecticides were fed to male and female rats for four generations, there was little variation in storage of the materials from one generation to another and no evidence of a continuing increase in succeeding generations (Adams et al., 1974). The concentrations of DDT in the blood and other tissues of the fetus are lower than those in corresponding tissues of the mother (Dedek and Schmidt, 1972). DDE, the dechlorometabolite of DDT, constitutes about 4% of technical DDT. Most species convert some of the DDT they ingest to DDE. Finally, most species, including humans, store DDE more tenaciously than they do DDT, the greater part of which is metabolized by a different pathway from that of DDE via DDA (Figure 93.2) and excreted more rapidly. The result is that DDE, expressed as a percentage of total DDT-related compounds, increases in individuals after DDT intake decreases and increases in successive steps of the food chain. The rhesus monkey apparently is an exception. Monkeys store DDE when it is fed to them. However, when feeding is stopped, the rate of loss of DDE stored in fat is more rapid than that of DDT (Durham et al., 1963). Whether it is relative inability to form DDE, unusual ability to excrete it, or a combination of both that accounts for the fact that little or no DDE can be found in monkeys fed DDT has not been entirely clear. The area under the concentration-time curve in plasma has been shown to be a more reliable predictor of liver response in rats than levels (Tomiyama et al., 2003, 2004).

93.2.3.5  Metabolism The metabolism of DDT is a complex affair and predominance of exact sequences are probably tissue- and speciesdependent as well as predomination of particular organs within an individual (Tebourbi et al., 2006). The chemical nature of the chief metabolite excreted in the urine was first elucidated by White and Sweeney (White and Sweeney, 1945). 2,2-bis(4-chlorophenyl) acetic acid (DDA) was isolated from the urine of rabbits chronically administered DDT. Later work by many authors confirmed that DDA isomers are the major urinary metabolites of p,p-DDT and o,p-DDT in all mammals, including humans, but the nature of all excreted metabolites and the relative importance of different routes between species still may not have been elucidated fully. The ability of phenobarbital and especially diphenyl­ hydantoin to promote the excretion of DDT was discovered in humans (Davies et al., 1969) and later confirmed in animals (Alary et al., 1971; Cranmer, 1970; Fries et al., 1971). This is of course consistent with our current knowledge of

Chapter | 93  Toxicology of DDT and Some Analogues

1983

CCl3

CCl2

Cl

Cl

p,p'-DDT

CHCl2

Cl

p,p'-DDD

Cl

Cl

Cl

p,p'-DDE CCl2

COCl *

HO Cl

Cl

Cl +

CCl2

Cl

Cl

Cl

+

CCl2

HO

HO

Cl p,p'-DDMU

Cl CH2Cl

CHO *

Cl HO

CHCl

CCl2

+

Cl

Cl

Cl

Cl

Cl

Cl COOH

Cl

CH2

CH2OH

Cl p,p'-DDA

Cl

Cl p,p'-DDOH

Cl

Cl p,p'-DDNU

Figure 93.2  Metabolites of p,p-DDT and the postulated route of metabolism in the rat. The metabolites indicated by an asterisk are likely to be reactive with cellular constituents.

the induction of drug metabolizing enzymes such as forms of cytochrome P450. That portion of the metabolism of DDT that leads to DDA in rats was explored by Peterson (Peterson and Robison, 1964) who produced evidence for the sequence of changes leading to DDA involving reduction to 1,1-dichloro-2,2-bis(4-chlorophenyl)ethane (DDD) followed by dehydrochlorination to 1-chloro-2,2-bis(4chlorophenyl)ethene (DDMU), which was apparently converted to 2,2-bis-(4-chlorophenyl)ethanol (DDOH) via 1-chloro-2,2-bis(4-chlorophenyl)ethane DDMS and 2,2-bis (4-chlorophenyl)ethane (DDNU) (see Figure 93.2). The compound identified as a “probable” intermediate aldehyde between p,p-DDA was later synthesized and shown to be highly labile (McKinney et al., 1969; Peterson and Robison, 1964) confirming that it is unlikely to accumulate in tissues in measurable amounts. Kujawa et al. obtained evidence for its formation from p,p-DDD by rat liver homogenates and its presence in the urine of rats injected with DDD (Kujawa et al., 1985). Two additional metabolites, bis(p-chlorophenyl)methane (DDM) and bis(p-chlorophenyl) methyl

ketone (DBP), were identified in chicks (Abou-Donia and Menzel, 1968). Not only was DBP found to result from the metabolism of DDA with DDM as an intermediate, but DBP was the only metabolite of DDE administered. Organ perfusion studies indicated that the liver is capable of biotransformation of DDT, DDE, DDD, DDMU, and other possible metabolites (Datta and Nelson, 1970). Cultures of human embryonic lung cells are capable of metabolizing DDT to DDA via DDD (North and Menzer, 1973). On the other hand, perfusion of rat, guinea pig, pig, and human skin samples has shown poor percutaneous passage of DDT (1%) and no evidence for cutaneous metabolism (Moir et al., 1994). When DDA was discovered, it was postulated that DDE was a step in its formation (White and Sweeney, 1945); however, rats which produced both DDE and DDA from DDT were said to be incapable of forming DDA when fed DDE (Peterson and Robison, 1964). This finding was contradicted by Datta (Datta and Nelson, 1970), who claimed that 14C-labeled p,p DDE was converted by rats to DDMU, which then underwent further metabolism to p,p-DDA.

1984

Datta suggested that the predominance of detoxication via DDE or DDD may depend on physiological response or the amount of toxicant used. DDE is stored in tissue as first demonstrated in connection with human fat (Mattson et al., 1953; Pearce et al., 1952). In fact DDE is stored more tenaciously than DDT. In vitro, the reductive dechlorination of p,p-DDT to DDD can occur with a cytochrome P450 system, especially under anaerobic conditions and induced by treatment in vivo with phenobarbital (Esaac and Matsumura, 1980; Hassall, 1971; Kitamura et al., 2002; Zaidi, 1987). A one-electron reduction of DDT to the 1,1-dichloro-2,2bis(p-chlorophenyl)ethyl radical seems to occur, followed by abstraction of a hydrogen atom, possibly from lipid, to give DDD (Kelner et al., 1986). The reduction of DDT to DDD is stimulated by thiols in an unknown manner. The formation of an intermediate radical explains binding to microsomal lipid, especially under anaerobic conditions (Baker and Van Dyke, 1984). DDD, on the other hand, needs aerobic conditions for binding, implying that further metabolism is required. Other studies with mouse liver microsomes have shown the formation of 2,2-bis(pchlorophenyl)-1,2-ethanediol (DDNU-diol) from DDNU, suggesting that a reactive epoxide intermediate might be formed (Planche et al., 1979), which might contribute to carcinogenicity. When synthesized, however, the ethylene oxide (DDNU-oxide) was not mutagenic. Gold and colleagues examined the metabolism of DDT metabolites in mice in vivo. The results seem to be a little different from that previously accepted for rats. It is thought that DDMU can undergo epoxidation; the resulting mutagenic epoxide is hydrolyzed and oxidized to 2-hydroxy-2,2-bis (4-chlorophenyl)acetic acid (OH-DDA), which is excreted in the urine (Gold et al., 1981; Gold and Brunk, 1982). Another route of metabolism of DDT in both the mouse and hamster seems to be the formation of DDA by a route involving hydroxylation on the C-1 side chain carbon of DDD (Gold and Brunk, 1982, 1983, 1984). Loss of HCl gives an intermediate acyl chloride, 2,2-bis(4-chlorophenyl)acetyl chloride (C1-DDA), capable of reacting with cellular proteins, DNA, etc., or losing water to give DDA. Since the work of Peterson (Peterson and Robison, 1964), the metabolism of DDT in rats was reexamined, with results similar to that described above for hamsters and mice (Fawcett et al., 1981, 1987). The conversion of p,pDDD to p,p-DDA occurs primarily by hydroxylation leading to C1-DDA, which on hydrolysis gives DDA. This acyl chloride may also be formed from DDE via an epoxidation route. Although DDMU is converted to DDA (Fawcett et al., 1987; Gold and Brunk, 1984), there is now considerable doubt as to whether it is an important intermediate in DDT metabolism. In addition, there is evidence to suggest that DDOH is a reduction product of DDCHO formed directly from DDD and not a precursor. A scheme for the metabolism of p,p-DDT in rats is shown in Figure 93.2 but

Hayes’ Handbook of Pesticide Toxicology

is still probably incomplete. Since metabolism stages occur in liver and kidney in vivo, routes may vary depending on the organ so that overall contributions are complex. The way in which DDE is lost from storage remained something of a mystery. In humans (Cueto and Biros, 1967), seals, and guillemots (Jansson et al., 1975), part of it is excreted unchanged, but the fact that its elimination is promoted by inducers of drug metabolism enzymes strongly suggests that much undergoes metabolism, conjugation, or both. That metabolism does occur was first demonstrated by identification of two hydroxylated derivatives of DDE in the feces of wild seals and guillemots and in the bile of seals (Jansson et al., 1975). When p,p-DDE was fed to rats, the same metabolites and one other were isolated from the feces, accounting for about 5% of the dose (Sundstrom et al., 1975). Later, a fourth hydroxylated derivative was identified in the feces of rats fed p,p-DDE. The metabolites are m-hydroxy-p,p-DDE [1,10-dichloro2-(p-chloro-m-hydroxyphenyl)-2,2(p-chlorophenyl)ethene the major metabolite], o-hydroxy-p,p-DDE, p-hydroxym,p-DDE (the product of an NIH shift), and p-hydroxyp-DDE. A scheme involving m,p-epoxy-p,p-DDE and o,m-epoxy-p,p-DDE was proposed for the formation of these metabolites as well as a fifth metabolite (Sundstrom, 1976). In mice, feeding DDE increased the hepatic levels of radioactivity from [14C]DDE and decreased that in the urine and feces (Gold and Brunk, 1986). The only metabol­ ite identified was the o-hydroxylated product. DDE is metabolized not only to easily excretable phenols but also to m-methylsulfone-p,p-DDE. In the blubber of seals from the Baltic, this compound was found in a concentration of 4 ppm along with DDE (138 ppm), DDD (10 ppm), DDT (78 ppm), and various polychlorinated biphenyls (PCBs) and their metabolites (150 ppm) (Jensen and Jansson, 1976). Sulfur-containing metabolites of halogenated aliphatic and aromatic chemicals usually arise by initial conjugation with glutathione followed by cleavage of individual amino acids, a lyase and methylation. The possibility of glutathione-derived conjugates of DDT requires further attention. Technical DDT contains the o,p-isomer. The interconversions of o,p-DDT and p,p DDT have been reported (Abou-Donia and Menzel, 1968; French and Jefferies, 1969; Klein et al., 1965), but there is considerable doubt as to whether these occur in vivo (Cranmer et al., 1972). Compared to p,p-DDT, the more rapid excretion of o,pDDT is explained at least in part by the observed ring hydroxylation of the parent compound in rats (Feil et al., 1973) and chickens (Feil et al., 1975) and of its metabolite o,p-DDD in rats (Reif and Sinsheimer, 1975) and humans (Reif et al., 1974) (see Figure 93.3). At least 13 metabolites were detected in rats. Ring hydroxylation, which has not been observed with p,p-DDT or p,p-DDD (but has been seen with p,p-DDE), occurs in all species but with species differences. For example, o,p-DDE and three hydroxylated

Chapter | 93  Toxicology of DDT and Some Analogues

1985

Cl

CCl3

Cl

CCl3

HO Cl

o,p'-DDT Cl

CHCl2

HO

Cl

Cl

CHCl2

o,p'-DDD Cl

Cl

Cl

Cl Cl

CHCl

COOH

Cl

Cl

CHCl2

o,p'-DDA

HO HO

Cl

Cl

Cl

COOH

COOH

HO

HO Cl

HO

Cl

3,4-OH o,p'-DDA Figure 93.3  Metabolism of o,p-DDT in the rat. In rats, glycine and serine conjugates of o,p-DDA have been found in the urine, and the aspartic acid conjugate of o,p-DDA has been found in the feces.

o,p-DDEs were found in the excreta of chickens but not in the excreta of rats. In two patients with adrenal carcinoma for which they were receiving o,p-DDD as an anti-tumor drug at a rate of 2000 mg/day, as much as 46–56% of the daily intake was recovered in the urine. Just over half of the recovered material was in the form of o,p-DDA, but the remainder was in the form of hydroxylated derivatives, mainly m-hydroxy-, p-hydroxy-, m-hydroxy-p-methoxy-, and p-hydroxy-m-methoxy-o,p-DDA. All hydroxylation had occurred on the ring that had its chlorine in the o position (Reif et al., 1974). When the metabolism of a single 100-mg oral dose of o,p-[14C]DDD was studied in rats, averages of 7.1 and 87.8% of the activity were recovered in the urine and feces, respectively, within 8 days (Reif and Sinsheimer, 1975). The high recovery indicated rapid excretion with little storage. o,p-DDD is specifically toxic for the adrenal cortex in a number of species including humans and is used as a drug (see in Section 93.3, TDE). In vitro studies suggest that this is due to its activation in adrenal mitochondria to a metabol­ ite which binds covalently. A metabolite is produced more polar than DDA, unlike the findings observed with the liver (Martz and Straw, 1977, 1980; Pohland and Counsell, 1985). 3-Methylsulfonyl-p,p-DDE is selectively covalently bound and toxic to the adrenal zona fasciculata of mice (Lund et al., 1988). A single dose of 3-methylsulfonyl-p,pDDE to mouse dams caused high binding in the adrenals of suckling pups with extensive vacuolation and necrosis

of the zona fasciculata. Slight degenerative changes were seen in fetal adrenals after dosing mothers with 50 mg/kg (Jonsson et al., 1992). The binding and damage probably result from cytochrome P450 activation (CYP11B) in adrenal mitochondria (Jonsson et al., 1991, 1995). A similar cytochrome P450-mediated activation may account for the covalent binding of o,p-DDD in mouse lung (Lund et al., 1986, 1989) and may be related to the acyl chloride formation already reported for p,p-DDT in rats and mice (see Figure 93.1). Cytochrome P450-mediated activation to the acyl chloride has also been proposed as partly accounting for the toxicity of DDD to isolated rabbit Clara cells and human bronchial epithelial cells (Nichols et al., 1995). Of the compounds shown in Figures 93.2 and 93.3, only DDT, DDD, DDE, and DDA commonly are reported in the tissues or excreta of animals, including humans. Conjugates of DDOH with fatty acids in the livers and spleens of rats given DDT have been reported (Leighty et al., 1980) and can be removed in vivo by treatment with bile salts, heparin, or lecithin (Leighty, 1981). Although microorganisms, plants, insects, and birds produce many of the same metabolites found in mammals, there are interesting differences. Nearly 20 derivatives (including mammalian metabolites) have been identified, and the chemical structures of others are still unknown. The metabol­ ism of microorganisms and plants, as well as that of domestic animals, may influence the composition of DDT-derived residues in human food, but there is no evidence that these

1986

residues contain a significant amount of any compound not formed from DDT by human metabolism.

93.2.3.6  Excretion When large doses of DDT are ingested, some of the compound is unabsorbed. Only traces of unaltered DDT may be found in the feces when exposure is by any route other than oral. However, true fecal excretion of DDT metabolites was established irrespective of the route of administration (Hayes, 1965), although either DDT metabolites are not excreted by humans in the feces to any important degree or they are excreted in one or more forms different from those demonstrated in rats. The bile appears to be the principal source of DDT metabolites in the feces of rats. When the bile duct was cannulated before intravenous injection of radioactive DDT, 65% of the dose was recovered in the bile, 2% in the urine, and only 0.3% in the feces (Jensen et al., 1957). The different routes of excretion are not unrelated. Burns et al. found that there was an increase in urinary excretion of radioactive material following ligation of the bile duct in rats fed radioactive DDT (Burns et al., 1957). This supports the finding by Jensen and his colleagues that most of the metabolites in bile are DDA or closely related to it. Although an enterohepatic circulation of the metabol­ ites of DDT has not been proved directly, it seems likely that such a circulation exists. The difference between the excretion of DDT and its metabolites in rats and the slower excretion in birds seems to be the reduced ability of birds to further metabolize DDE (Fawcett et al., 1981). The excretion of DDE in rats is dependent on dose and probably involves induction of drug-metabolizing systems (Ando, 1982). The excretion of DDT in milk was first published by Woodard et al. (1945) in connection with a dog fed at the rate of 80 mg/kg/day. Within a short time, excretion of DDT in milk was reported in rats, goats, and cows, and in 1951 it was demonstrated in women (Laug et al., 1951). Rats fed a diet containing 1000 ppm produced milk that was toxic to their young (Telford and Guthrie, 1945). Following these early studies, the presence of DDT was demonstrated repeatedly in the milk of cows. Cows fed substantial, not nontoxic, residues of DDT commonly excrete 10% or more of the total dose in their milk (Hayes, 1959). The proportion of the mother’s DDT intake that could be recovered from her milk varied from 12.6 to 30.2% and averaged 24.6% in rats receiving the compound from their diet at an average rate of 32.4 mg/kg/day. Under these circumstances, the dosage of the young was somewhat less than half of that of their mothers on a milligram per kilogram basis. The oral dosage of 32.4 mg/kg/day was well tolerated by both dams and pups, as was also true of an intraperitoneal dosage of 100 mg/kg/day. An intraperitoneal dosage of 200 mg/kg/day killed some dams, but most of the pups of other dams survived. All of the pups of these mothers experienced reduced milk intake and reduced weight gain.

Hayes’ Handbook of Pesticide Toxicology

The concentration of DDT in the brains of these pups was much lower than in pups killed by oral administration of the compound, indicating that the young of mothers receiving massive dosages of DDT suffer malnutrition but not poisoning (Hayes, 1976). Wilson et al. showed that DDT was secreted from the skin of a cow maintained on an oral dosage of about 53 mg/kg/day (Wilson et al., 1946). Because DDA is the main form in which DDT is excreted, it might be expected that, following its direct administration, DDA would be excreted relatively efficiently. During the first several days after oral dosing, rabbits excreted DDA in the urine approximately 15 times faster than animals given DDT at an equivalent dosage. Although the rate of DDA excretion associated with DDT increased more rapidly, so that the values differed by a factor of only 5 after day 20 of feeding (Smith et al., 1946).

93.2.3.7  Biochemical Effects The main mechanism of action of DDT is its effect on membranes in the nervous system, especially axonal membranes. The effect on axons may be related to inhibition of Na-, K-, and Mg2-adenosine triphosphatase derived from a nerve ending fraction of rabbit brain that is inhibited by DDT. A similar enzyme that binds DDT was isolated from the synapses of rat brain (Bratowski and Matsumura, 1972). There has been considerable interest in a Ca-ATPase, which may regulate calcium levels at the axon surface (Ghiasuddin and Matsumura, 1979). In insects, a protein subunit of ATP synthase has been proposed as the DDT-sensitive target (Younis et al., 2002). DDT is known to cause prolonged opening of the ion gates of the sodium channel perhaps by affecting phosphorylation in the -subunit protein (Ishikawa et al., 1989). Song et al. showed that DDT appeared to interact with sodium channels of rat dorsal, not ganglion neurons, in the same manner as type I and type II pyrethroids (Song et al., 1996). At a supralethal dosage of 600 mg/kg, DDT caused in rats a marked decrease in the concentration of cortical and striatal acetylcholine and of brain stem 3-methoxy-6-hydroxphenylglycol and 5-hydroxyindoleacetic acid (Hrdina et al., 1973; Hudson et al., 1985; Tilson et al., 1986). p-Chlorophenylalanine blocked all of the neurotoxic signs of poisoning, and other inhibitors blocked one or another but not all of the effects. It was suggested that changes in the metabolism of 5-hydroxytryptamine and norepinephrine may be responsible for DDT-induced hypothermia and acetylcholine may be related to tremors and convulsions (Hrdina et al., 1973). Although spinal -adrenoceptors have been proposed as modulating DDT-induced tremor (Herr and Tilson, 1987), attenuation of DDT-induced motor dysfunction requires blockade of -adrenoceptors in regions other than solely the spinal cord (Herr et al., 1989). At a lower dose of DDT (180 mg/kg), but one which still induced convulsive tremor,

Chapter | 93  Toxicology of DDT and Some Analogues

acetylcholine and cyclic GMP were increased in the cere­ bellum (Aldridge et al., 1978). In adult rats and mice, there is a decrease in the cholinergic muscarinic receptors of rat brain (Eriksson et al., 1984), particularly in the cerebellum (Fonseca et al., 1986). The palmitic acid conjugate of DDOH can also have this effect (Eriksson et al., 1984). Disturbances of brain lipid metabolism have been observed in monkeys after chronic exposure to DDT (Sanyal et al., 1986). Khaikina and Shilina reported that administration of DDT to rats at only one-fifth of the LD50 for 20 days increased by 188% the amount of 5-hydroxyindoleacetic acid excreted in their urine (Khaikina and Shilina, 1971). This indicated a change in the metabolism of serotonin, but it probably does not support a serotonin deficiency as a DDT mode of action (Chung Hwang and Van Woert, 1981). It is evident that many of the side effects of DDT are the result of its induction of drug metabolizing enzymes. Oral administration of o,p-DDT to dogs at a rate of 50 mg/kg/day stimulates the microsomal enzymes of the liver. These changes in the liver are initially accompanied by an increase in the size of the adrenals and of the cells of the zona fasciculata; these cells become vacuolated and devoid of acidophilic cytoplasm, and their nuclei become hyperchromatic and often peripheral in position. Synthesis of corticosteroids by the adrenal is not blocked (Copeland and Cranmer, 1974). Thus the effect of a substantial dosage of o,p-DDT is quite different from that of o,p-DDD, although part of the metabolism of o,p-DDT must be by that route. The tissue level of p,p-DDE necessary to induce liver microsomal enzymes is lower in the rat than in the quail (and possibly in other birds). Thus Bunyan et al., using residues in the heart as an index, found a maximal increase in cytochrome P450 per weight of liver and a maximal activity of aniline hydroxylase activity at tissue levels of approximately 3 ppm DDE in rats and 40 ppm DDE in quail (Bunyan et al., 1972). However, at any given dietary level, higher tissue levels were reached by quail than by rats, so the dosage responses of the two were similar. These authors concluded that DDE is more important than DDT in inducing microsomal enzymes, but in humans the opposite appears to be true. The significance of the induction of hepatic enzymes and any correlation with the potential hepatocarcinogenicity of DDT can be found elsewhere in this chapter and is particularly discussed in Smith (1991). In the female rat, multiple bi-daily doses of DDT induced hepatic CYP2B and 3A proteins but not CYP1A1 or 1E1 (Li et al., 1995) and caused elevated hydroxylation at positions 16 and 6 of testosterone. DDT, DDE, and DDD all induced CYP2B and 3A in male rat liver to not dissimilar degrees despite marked differences in bioretention (Nims et al., 1998). Induction of CYP2B and 3A expression in rat liver by DDE is mediated by the constitutive androstane and pregnane X receptors (Wyde et al., 2003).

1987

In squirrel monkeys (and presumably in other species) only 2 days on a vitamin C-deficient diet impairs both the induction of o-demethylase and the stimulation of the glucuronic acid system by DDT (5 mg/monkey/day) (Chadwick et al., 1971). In guinea pigs, maintenance of induction of microsomal enzymes requires a higher dietary level of vitamin C than does prevention of scurvy (Wagstaff and Street, 1971). Since lipids are associated with the function of microsomal enzymes and DDT induces these enzymes, it might have been expected that DDT and essential fatty acids would interact. Tinsley and Lowry (1972) found that the growth of female rats receiving p,p-DDT at a dietary level of 150 ppm was depressed if they received a diet deficient in essential fatty acids but was slightly stimulated if they received the same diet supplemented with these acids. It was suggested that DDT influenced essential fatty acid metabolism by increasing the demand for them. Sampson et al. found that DDT did not exacerbate aspects of essential fatty acid deficiency but did alter lipid metabolism in an unexplained way (Sampson et al., 1980). Exposure of rats to DDT by the intratracheal route has shown lung lipid metabolic changes but the significance is unclear (Narayan et al., 1990a,b). In contrast, a variety of diets (containing fats that may occur in the human diet and that were in approximately the same proportion as fats in typical human food in the United States) had little or no influence on the storage of DDT and a wide range of pesticides fed to rats for four generations in combination at rates only 200 times those found in food in the United States (Adams et al., 1974). Fat mobilization can cause rapid release of stored DDT, but this does not seem to be associated with any major toxic effect assessed pathologically or biochemically (Mitjavila et al., 1981). DDT has been shown in vitro and sometimes in vivo to influence some enzymes of intermediary metabolism and other miscellaneous enzymes. For instance, DDT and a variety of analogues have been shown to affect isolated rat liver mitochondria, but the significance of this in vivo is uncertain (Ohyama et al., 1982). Whether this is linked to the effects caused by 3-methylsulfonyl-p,p-DDE in adrenal mitochondria is not known. The hyperglycemia observed during much of the early part of acute poisoning may be associated with an increase in four gluconeogenic enzymes (pyruvate carboxylase, phosph­oenolpyruvate carboxykinase, fructose-1,6diphosphatase, and glucose-6-phosphatase) (Kacew and Singhal, 1973). Increase in these enzymes in the renal cortex of rats has been observed after a single dose at a rate as low as 100 mg/kg or greater or following 45 daily doses at rates of 5 or 25 mg/kg/day. The changes are not mediated through release of corticosteroids from the adrenal glands. The fact that 100 mg/kg is the smallest single dosage that produced a statistically significant change in these enzymes indicates that their alteration is a complication rather than a cause of poisoning. High concentrations of DDT inhibit phosphatidase, muscle phosphatases, carbonic anhydrase,

1988

and oxaloacetic carboxylase and increase the activity of cytochrome oxidase and succinic dehydrogenase. However, none of these changes, with the possible exception of inhibition of carbonic anhydrase, may have any connection with the toxic action of DDT or even with its side effects. To the author’s knowledge, there have been no full toxicogenomic studies of DDT, which would probably reveal that many changes in gene expression are associated with general mechanisms of toxicity. Neal et al. reported a small but consistent increase in the volume of urine excreted in 24 h when dogs were dosed orally or by insufflation at the rate of 100 mg/ kg/day (Neal et al., 1944). The possibility that increased urinary output is related to the inhibition of carbonic anhydrase (Torda and Wolff, 1949) may deserve attention, but data from volunteers receiving 3.5 or 35 mg/person/day indicated no increase in urinary volume compared with controls (Hayes et al., 1971). Many enzymes including plasma amylase, aldylase, glutamic pyruvic transaminase, and isocitric dehydrogenase were not changed in squirrel monkeys given dosages from 0.05 to 50 mg/kg/day, the latter of which proved fatal within 14 weeks (Cranmer et al., 1972).

93.2.3.8  Effects on the Nervous System The major toxic action of DDT is clearly on the nervous system, probably by slowing down closing of “gates” in axon sodium channels (Dubois and Bergman, 1977; Hong et al., 1986; Woolley, 1985), and it requires an intact organism for full expression. For other biochemical mechanisms related to the nervous system, see Section 93.2.3.7. The fact that DDT causes a myotonic response in muscle and substitution of a train of spikes for the normal diphasic electroneurogram (Eyzaguirre and Lilienthal, 1949) is in marked contrast to the absence of detectable injury or, in fact, any response in other isolated tissues. In spite of the importance of the nervous system, a detailed review of early literature indicates that although the presence of some specialized nervous function may be necessary for the manifestation of DDT poisoning, the mere occurrence of specialized nervous fibers in certain protozoa or the occurrence of a rather complex nervous system in mollusks is not sufficient to render these forms susceptible. Just as there is little explanation still for the effect of DDT in susceptible species, the fact that certain species and even entire phyla are inherently resistant to the compound is still not entirely understood. Probably all major parts of the nervous system, both central and peripheral, are affected by DDT. Whereas effects on specific portions, notably the cerebellum and the motor cortex, have been viewed as of greatest importance, it probably is more accurate to emphasize the interaction of functions, all modified to some degree. Farkas et al. found that electrocardiogram wave frequency showed considerable increase in resting rats that had received 20 mg/kg/day as a result of dietary intake.

Hayes’ Handbook of Pesticide Toxicology

Rats that had received 5 mg/kg/day did not exhibit this change while at rest, but even these exhibited abnormalities when exposed to a rhythmic light stimulus (Farkas et al., 1968). Electrical activity may become abnormal only 1 or 2 min after administration of a large dose of DDT. Four stages culminating in generalized seizure were described by Joy (1973). Phenobarbital, but not diphenylhydantoin or trimethadione, was effective in stopping seizures. The most characteristic effect of DDT in contrast to dieldrin, for example, is the production of tremor. Sufficient dosages of DDT produce tremor even at ambient temperatures that approach body heat. However, dosages of DDT that produce no other clinical effect make rats more sensitive to low temperatures. This sensitivity may be demonstrated by having the rats swim to exhaustion in cool water. The ability of the rat to keep afloat is more dependent on coordination than on physical strength. DDT appreciably reduces the swimming time (Smith, 1991). Like tremor, the coldness of the skin and ruffling of the fur seen in acute poisoning probably represent an indication of disturbed thermal regulation. Apparently, it was not until the work of Hrdina et al. that a change of almost 3°C in body temperature was reported in rats following a fatal (600 mg/kg) oral dosage of DDT (Hrdina et al., 1975). The central nervous systems of mice and hamsters are equally sensitive, the concentration of DDT in their brains at death being similar (Gingell and Wallcave, 1974). However, after an oral dosage of 500 mg/kg, the DDT concentration of the mouse brain was twice that of the hamster. This cannot be explained by a difference in absorption, metabolism, or excretion but apparently is due to a difference in permeability of the blood–brain barriers of the two species. When animals received DDT at a dietary level of 205 ppm for 6 weeks, the residues in fat and liver were seven to eight times higher in the mouse, a fact only partially explained by the greater food intake of mice relative to body weight. Although urinary excretion of [14C] DDT was similar in previously unexposed hamsters and mice, this excretion was stimulated in the hamster but little affected in the mouse by previous dietary exposure to DDT. Careful studies have shown that neurotoxic actions of DDT following oral dosing of rats are significantly affected by the volume of the dosing solution. Not surprisingly, this is probably the consequence of higher partitioning of DDT in oil and greater gut motility, but it illustrates the importance of pharmacokinetic knowledge in comparative studies (McDaniel and Moser, 1997). Both DDT and DDE interact similarly with model and native membranes, causing disordering effects in cholesterol-rich membranes, such as brain microsomes (Antunes-Madeira Mdo and Madeira, 1993). Whether these effects have any bearing on neurotoxicity in vivo is unknown. Specific depletion of dopaminergic neurons in primary cultures from rats by DDT has been reported (Leung et al., 2003).

Chapter | 93  Toxicology of DDT and Some Analogues

93.2.3.9  Cause of Death Death from DDT poisoning is usually the result of respiratory arrest. The heart continues to beat to the end and in some instances continues a little while after respiration stops. Deichmann et al. (1950) found that the onset of hyperirritability was accompanied by an increase in the frequency and amplitude of respiration (quoted in Hayes, 1982). Later, with the occurrence of tremors, the depth of respiration frequently returned to a more normal level, but the rate remained high. In some animals, respiration stopped suddenly after a deep inspiration during a tonic convulsion. In other animals, the rate and amplitude decreased progressively and finally ceased without any terminal spasm. Animals that die of respiratory failure caused by DDT do so after a relatively long period of muscular activity that leaves them exhausted. It was shown by Phillips and coworkers that the hearts of dogs given large intravenous doses of DDT were sensitized to epinephrine (Phillips and Gilman, 1946; Phillips et al., 1946). This was true not only of injected epinephrine but also of the compound released by the adrenal glands during a seizure. Stimulated in this way, the sensitized hearts of dogs developed an irreversible, fatal ventricular fibrillation. However, the hearts of monkeys were able to recover from fibrillation and resume normal rhythm. It is not clear how important sensitization of the myocardium is when DDT is administered by other routes, but ventricular fibrillation may be the cause of death in animals that die suddenly soon after onset of poisoning. Thus, DDT not only sensitizes the myocardium in a way similar to that of halogenated hydrocarbon solvents but also, through its action on the central nervous system, produces the stimulus that increases the likelihood of fibrillation. There is no evidence that repeated, tolerated doses of DDT sensitize the heart (Jeyaratnam and Forshaw, 1974). Rats were fed DDT at a dietary level of 200 ppm (about 10 mg/kg/day) for 8 months, during which they received weekly intraperitoneal doses of vasopressin, which causes a temporary myocardial ischemia. Electrocardiograms showed no significant increase in cardiac arrhythmias in the DDTfed rats compared with controls. Intravenous noradrenaline given at the end of the 8-month period did not produce a greater incidence of arrhythmias in the DDT-fed rats.

93.2.3.10  Mutation and Carcinogenesis DDT has been tested in a number of ways for possible mutational effect. Much of this work has been reviewed in detail together with most of the carcinogenicity studies shown in Table 93.3 (Coulston, 1985a,b). For example, DDT was listed as a negative chemical in microbial mutagenicity screening studies including metabolic activation systems with both DDT and DDE (McCann and Ames, 1976; Shirasu et al., 1976).

1989

At a dosage of 105 mg/kg DDT, produced no increase of dominant lethals in mice (Epstein and Shafner, 1968). However, concentrations of 10 ppm or greater produced chromosome breaks and exchange figures in a marsupial somatic cell line (Palmer et al., 1972). A slight mutagenic effect in mammals has been reported by Markarian (1966). Deletions plus gaps were reported to be more common in the chromosomes of mice that had received DDT. On the whole, in vitro tests of the mutagenicity of DDT have given only negative or dubious results (Coulston, 1985a,b). An unconventional test for mutagenicity involved examination of explants of pulmonary tissue from embryonic mice whose dams had been fed dietary concentrations of 10 and 50 ppm DDT (Shabad et al., 1972). An increase of diffuse hyperplasia and focal proliferation was observed, but a dosage– response relationship was not clear. Some of the embryos were allowed to live and the experiment was repeated in subsequent generations. There was no continuing progression of the reported changes in succeeding generations. The question of whether DDT is carcinogenic really seems to be restricted, experimentally, to its action in the liver of some rodents. Some of the positive findings shown in Table 93.3 have not been found in other studies (NCI, 1978a). However, there is still the evidence that DDT can act as a promoter of liver carcinogenesis initiated by aflatoxin and of other chemicals in vitro and in vivo (Peraino et al., 1975; Rojanapo et al., 1987; Schulte-Hermann, 1985; Sugie et al., 1987; Williams and Numoto, 1984). DDT causes inhibition of intercellular communication in cultured rat liver epithelial cells and in hamster cell lines (Flodstrom et al., 1990; Tsushimoto et al., 1983; Warngard et al., 1989; Williams et al., 1981). This could be protected against by extracts of green tea (Sigler and Ruch, 1993). Using freeze-fracture analysis of hepatocytes from rats exposed in vivo to DDT, both gap junction size and number were reduced (Sugie et al., 1987); changes in connexin 32 and connexin 26 expression could be responsible (Tateno et al., 1994). Additional studies with the rat “oval” cell line WB-F344, which does not express connexin 32 but rather connexin 43 predominantly, showed not a decreased expression but a decrease in the phosphorylated form (Ruch et al., 1994). Subsequent investigations suggest this is due to endocytosis of gap junctions and degradation of lysosomal connexin 43 P2 (Guan and Ruch, 1996). The tumorigenic action of DDT on the liver of rodents is usually considered to be like that of phenobarbital. Recent studies demonstrate that the mode of action of phenobarbital and such chemicals in rodents, including changes in gene methylation status, are probably not relevant to humans (Holsapple et al., 2006). Earlier discussion on the effects of DDT in relation to hepatocarcinogenicity can be found in a previous edition of this chapter. Some evidence suggests that o,p-DDT can support the growth of an estrogen-responsive tumor (Robison et al., 1985) and there have been vigorous debates on the role of

1990

DDT and analogues in human breast and testis tumors (see Section 93.2.4). DDT has been used as a model compound to demonstrate estrogen-independent stimulation of the P38 MAPK pathway leading to phosphorylation of p300, a general transcription factor essential in growth, differentiation, and cell death (Bratton et al., 2009; Frigo et al., 2002, 2004).

93.2.3.11  Other Miscellaneous Effects on Organs and Tissues Many early reports reviewed by Hayes indicate that large doses of DDT may have no effect on the blood or they may produce a moderate leukocytosis and a decrease in hemoglobin, with or without a decrease in the concentration of red cells (Hayes, 1959). The leukocytosis probably is secondary to stimulation of the sympathetic nervous system, while the loss of hemoglobin may be nutritional in origin. A later study with squirrel monkeys did not confirm the early results (Cranmer et al., 1972). A range of hematologic parameters remained unchanged in squirrel monkeys dosed orally at rates of 0, 0.05, 0.5, 5, and 50 mg/kg/day, even though the highest dosage was fatal within 14 weeks. Average protein-bound iodine levels of 5.42 and 6.93 g/%, respectively, were reported in the sera of 42 workers occupationally exposed to chlorinated hydrocarbon insecticides and 51 workers not so exposed. The differences were statistically significant even though all values fell within the normal range of 4–8 g/% (Wassermann et al., 1971). It was not recorded whether the workers involved were from the same factory as those with 10 or more years of occupational exposure whose plasma DDT levels were previously reported. After a single large dose (100 mg/kg) to rats, thyroidal 131I release was completely inhibited for more than 12 h (Goldman, 1981). The view of Clifford and Weil was that there was no evidence that occupational exposure to DDT has had any effect on human endocrine organs (Clifford and Weil, 1972). What at first appeared to be an immunological response to DDT in guinea pigs really involved a quite different, predictable effect. Animals sensitized to diphtheria toxoid were less susceptible to anaphylaxis in response to a challenge dose of the toxoid if they were pretreated with DDT at a dosage of only 10–20 mg/kg/day (Gabliks et al., 1973, 1975). Direct measurement of antitoxin production indicated little or no difference between protected and unprotected animals. Furthermore, some protection was given by DDT administered for only 3 days prior to the induction of anaphylaxis. Further study showed that DDT treatment reduced the histamine levels in the lungs of both immunized and nonimmunized animals. The number of detectable mast cells was also reduced; this was true whether the count was made in tissues from guinea pigs dosed systemically with DDT or in lungs and mesenteries from untreated animals exposed to DDT in vitro at concentrations ranging from 10 to 45 ppm. These results indicated that the

Hayes’ Handbook of Pesticide Toxicology

protection offered by DDT was the result of a reduction of the amount of histamine available for sudden release in response to a challenge dose of toxoid (Askari and Gabliks, 1973). Regardless of exposure to DDT, immunization leads to an increase in detectable mast cells (Gabliks et al., 1975). Banerjee (Banerjee et al., 1996) compared DDT with some of its metabolites in suppressing aspects of humoral and cellular immune response in rats: DDE and DDD but not DDA appeared to play a role. DDT depresses activation of mouse macrophages and decreases resistance to infection by some bacteria (Nunez et al., 2002). Dutta and coworkers consider that functional and structural effects on mouse macrophages, together with excessive and uncontrolled complement consumption are major mechanisms in the immunosuppression effects of DDT (Dutta et al., 2008). DDT has been reported to cause acute renal failure in rats after intravenous infusion (Koschier et al., 1980).

93.2.3.12  Effects on Reproduction In recent years, the possibility that environmental levels of DDT isomers and metabolites might have effects on human reproductive function and carcinogenesis in breast and testes has become a major issue. It was shown very early that DDT produces a striking inhibition of testicular growth and secondary sexual characteristics of cockerels when injected subcutaneously in dosages as high as 300 mg/kg/day (Burlington and Lindeman, 1950). Changes in the testis involve the tubules and not the interstitial tissues, and they have been attributed to an estrogen-like action of DDT. Before the problem of residues became evident, DDT was used extensively for control of lice and common mites on chickens without any adverse effects on egg production or other aspects of reproduction. It must be emphasized that many rats would be killed the first day if they were given the dosage of DDT that has been shown to affect the testis of cockerels. The report that under special conditions DDT has a gonadotoxic effect is of questionable significance in view of the results of multigeneration tests in rats, mice, and dogs (Rybakova, 1968). Dean et al. were unable to demonstrate any changes in either serum androgens or testicular synthesis of testosterone in young rats after exposure to DDT, despite significant induction of metabolism of testosterone by isolated hepatic microsomes (Dean et al., 1980), although changes in testosterone metabolites have been observed in vivo (Sierra-Santoyo et al., 2005). However, more recent evidence proposes that DDT and its metabolite DDE can disrupt male reproductive development and act as antiandrogens by binding androgen receptors in a nonproductive manner (Kelce et al., 1995, 1998; Sakly, 2001). DDE, like vinclozolin and flutamide, changed the expression of androgen receptor-regulated genes, such as the prostatic mRNA prostatein C3, in castratedtestosterone-treated male rats (Kelce et al., 1997). Exposure of male rats to DDE in utero and by lactation showed that

Chapter | 93  Toxicology of DDT and Some Analogues

by some parameters Long-Evans rats were more sensitive than Sprague–Dawley, e.g., ano-genital distance, but DDE had no effects on testes, epididymis, seminal vesicles, or ventral prostate weights (You et al., 1998). Effects were minimal at maternal doses less than 10 mg/kg/day. Confirmation that Long-Evans rats were more sensitive to DDE than some other strains was reported from studies designed to detect antiandrogen endocrine disrupters and appeared to act centrally rather than peripherally (O’Connor et al., 1999). The effects of p,p-DDE on the development of the rat prostate was confirmed with Holtzman rats exposed in utero or by lactation (Loeffler and Peterson, 1999) without changes in serum androgen concentrations. This suggests effects on androgen signaling pathways within the prostate. Some evidence for a partial interaction with the antiandrogenic effects of TCDD was suggested from dual dosing experiments. The significance of these studies with regard to human male development is really not clear given the considerably greater experimental exposures. DDT is able to activate estrogen receptors in brain and liver but probably following metabolism (Mussi et al., 2005). Intraperitoneal injection of as little as 5 mg/kg of technical DDT or 1 mg/kg of o,p-DDT causes a significant increase in weight of the uterus of normal immature female rats or of ovariectomized adult females (Welch et al., 1969). A very much smaller stimulation is caused by p,p-DDT. Treatment of rats with DDT, especially o,pDDT, inhibited uptake of estrogen by the uterus in vivo, possibly by competition for binding sites. Isomers of DDD and DDE do not influence uterine weight or the binding of estradiol (Welch et al., 1969). It seems unlikely that metabolic activities of o,p-DDT is necessary as is true of activation of o,p-methoxychlor (Kupfer and Bulger, 1979). The action of o,p-DDT on the uterus seems to be as a long-acting agonistic estrogen interacting with the same receptor as 17-estradiol (Galand et al., 1987; Ireland et al., 1980; Robison et al., 1984). However, some differences from estradiol have been recorded (Robison et al., 1985). The lesser enantiomer of o,p-DDT seems to be the active isomer (McBlain, 1987). The binding and estrogenic activity of DDT analogues in rats is only about 1/10,000 as great as that of diethylstilbestrol (Nelson, 1973), an important point when considering potential effects of trace levels relative to endogenous estrogens. o,p-DDT inhibited DNA synthesis in cultured bovine oviductal and uterine cells to a greater degree than methoxychlor, especially uterine epithelial and stromal cells (Tiemann et al., 1996). However, at lower concentrations both o,p-DDT and methoxychlor stimulated DNA synthesis. A considerably smaller dosage of o,p-DDT resulting from a dietary level of 10 ppm for 2–9 months had no effect on reproduction in ewes (Wrenn et al., 1971b). In a similar way, dietary levels of o,p-DDT as high as 40 ppm, giving a dosage level of about 2.1 mg/kg/day in rats, failed to interfere with reproduction and lactation in these animals,

1991

although dosage was continued through two pregnancies (Wrenn et al., 1971a). The report (Heinrich et al., 1971) that o,p-DDT significantly advances puberty, induces persistent vaginal estrus after a period of normal estrus cycles, and causes other reproductive abnormalities in female rats would at first appear inconsistent with the lack of effect of technical DDT or o,p-DDT on reproduction cited above. The same is true of other effects of o,p-DDT subcutaneously on the 2nd, 3rd, and 4th days of life (counting the day of birth as zero). Because rat pups on the 3rd day weighed about 12 g or less each, it follows that the subcutaneous dosage was about 83.3 mg/kg/day or more, that is, about 40 times greater than the highest oral dosage of o,p-isomer fed to breeding rats and more than 105 times greater than human dietary exposure. Ottoboni found that female rats reproduced normally when fed DDT for two generations at dietary levels as high as 200 ppm (about 10 mg/kg/day except during lactation, when intake is increased about threefold) (Ottoboni, 1969). In fact, at a dietary level of 20 ppm, the dams had a significantly longer reproductive life span (14.5 months) than their littermate controls (8.9 months); the number of females becoming pregnant after the age of 17 months and the number of successful pregnancies after that age were significantly different in the two groups (Ottoboni, 1972). In a more recent study, changes in serum levels of estradiol and progesterone and/or a delay in male rats’ sexual maturation was noted in rats fed DDT at 50 and 350 ppm in the diet, but except for a decrease in pup viability at the highest dose, there were no substantial reproductive disorders (Hojo et al., 2006). In a study focused mainly on DDT in milk, the full ability of rats to reproduce at a dietary level of 200 ppm was confirmed, and the ability of dams injected intraperitoneally at levels as high as 100 mg/kg/day to rear their young was demonstrated (Hayes, 1976). A six-generation test of reproduction in mice showed no effect of DDT at a dietary level of 25 ppm on fertility, gestation, viability, lactation, and survival (Keplinger et al., 1970). A level of 100 ppm produced a slight reduction in lactation, and survival in some generations but not all, and the effect was not progressive. A level of 250 ppm was distinctly injurious to reproduction. The dietary concentrations used determine dosages of 3.33, 13.3, and 33.2 mg/kg/day in nonpregnant, nonlactating, adult female mice. The intake is much higher in both young and lactating mice (see Smith, 2001 for reference). Four female dogs of unstated age that previously had received DDT at the rate of 12 mg/kg/day, 5 days/week, for 14 months were bred when they went into heat (Deichmann et al., 1971). The males involved had been fed aldrin (0.15 mg/kg/day) plus DDT (60 mg/kg/day) for 14 months prior to breeding but not during breeding. Two of the females went into heat but failed to become pregnant, and

1992

one failed to come into heat during 12 months after feeding stopped. Four of six pups born to the fourth female died within 1 week of birth; the other two were weaned successfully even though only two posterior mammae of the mother were functional. A three-generation study failed to confirm any of the injuries suggested by the study of four dogs. In the three-generation study, male and female dogs were fed technical DDT from weaning at rates of 0, 1, 5, and 10 mg/ kg day. Observations were made on 135 adult females, 63 adult males, and 650 pups. There were no statistically significant differences among controls and DDT-treated dogs in length of gestation, fertility, success of pregnancy, litter size, or lactation ability of the dams; in viability at birth, survival to weaning, sex distribution, and growth of pups; or in morbidity, mortality, organ/body weight ratios, or gross histological abnormalities in all the animals studied. The only clear difference was that DDT-treated females had their first estrus 2 or 3 months earlier than the control dogs. There was a slight increase in liver/body weight ratio in some DDT-treated animals but the difference was not statistically significant, not dosage related, and not associated with any histological change (Ottoboni et al., 1977). When p,p-DDT was administered to pregnant mice at a rate of 1 mg/kg on days 10, 12, and 17 of gestation, it was not teratogenic but did alter the gonads and decrease the fertility of the young, especially the females (McLachlan and Dixon, 1972). A single dose at the rate of 15 mg/kg or repeated doses of 2.5 mg/kg/day given during pregnancy may be embryotoxic but not teratogenic to mice (Schmidt, 1973). DDT was shown to be more toxic than methoxychlor to preimplantation mouse embryos in culture (Alm et al., 1996). Teratogenic effects of DDT have not been seen in studies of reproduction, including those for two generations in rats, six generations in mice, and three generations in dogs (Smith, 1991). Because of the estrogenic properties of large doses of DDT, the compound was considered as a possible cause of abortion in dairy cattle, but no evidence for a relationship was found (Macklin and Ribelin, 1971). A similar conclusion was reached regarding human abortions (O’Leary et al., 1970).

93.2.3.13  Behavioral Effects Behavioral changes may be demonstrated in animals receiving DDT daily at rates too low to produce illness. Khairy (Khairy, 1959) detected ataxia in the form of changes in gait in rats that had been fed DDT at dietary levels of 100 ppm or more for 21 days. The results were recorded in terms of the tangent, that is, the ratio of the width and length of step. At a dosage of about 5 mg/kg/day, the ratio was less than normal, a change attributed to an exaggeration of the stretch reflex. At dosages of about 10, 20, and 30 mg/kg/day, the ratio was progressively increased above normal as a result of broadening of the gait and shortening of the steps. These same dosage levels did not affect problem-solving behavior or speed of locomotion. The

Hayes’ Handbook of Pesticide Toxicology

experimental animals were found to be generally less reactive to stress than normal ones. The acoustic startle response of rats is significantly increased after a 12.5 mg/kg dose of p,p-DDT but can be attenuated by phenytoin and an adrenergic receptor antagonist, phenoxybenzamine (Herr et al., 1987; Herr and Tilson, 1987; Saito et al., 1986; Tilson et al., 1985), which also decreased DDT-induced myoclonus (Hwang and Van Woert, 1978). See also Section 93.2.3.8.

93.2.3.14  Pathology Morphological changes are inadequate to account for death from DDT poisoning. Changes that occur in the liver have been discussed previously (Smith, 1991). Mild to moderate morphological changes have been reported in the kidneys of animals that had received massive single doses or repeated doses; examples are fatty degeneration, necrosis, and calcification (Lillie et al., 1947; Stohlman and Lillie, 1948) or slight brown pigmentation of the convoluted tubular epithelium (Fitzhugh and Nelson, 1947). However, it sometimes has happened that a complete absence of change in the kidney has been reported in connection with other studies carried out in the same laboratories (Lillie and Smith, 1944; Nelson et al., 1944).

93.2.3.15  Treatment of Poisoning in Animals The more successful studies of treatment of animals poisoned by DDT involve the nervous system. A full discussion of this can be found in Smith (1991). In essence, sodium phenobarbitol affords little help in rats but in dogs and cats, and to a lesser extent monkeys, it protected against tremors and death (Phillips and Gilman, 1946). Later reports showed that phenytoin attenuates the tremor produced in rats by DDT and permethrin but not by lindane and chlordecone (Tilson et al., 1985, 1986). Vaz and his colleagues (Vaz et al., 1945) were apparently the first to note the antidotal effect of calcium in DDT poisoning. Dogs were given DDT orally as a 10% oily solution at a daily dosage of 100 mg/kg until signs of intoxication appeared. The same dosage could then be repeated to produce intense symptomatology from which the animals would recover spontaneously in 12–24 h. For the actual tests, a larger challenge dosage of DDT (150–200 mg/kg) was used. Each dose of calcium gluconate (30 ml of a 10% solution) was injected intravenously into dogs weighing 8–18 kg. Dogs that were injected with calcium gluconate daily for 4 days and challenged with a large dose of DDT on the 4th day developed no symptoms or only slight ones. Dogs receiving a single dose of calcium gluconate showed symptoms of short duration and survived following a dosage of DDT large enough to kill two controls. Koster studied cats poisoned by the intravenous injection of a soya lecithin–corn oil emulsion of DDT (Koster, 1947). A comparison was made of several aspects of intoxication,

Chapter | 93  Toxicology of DDT and Some Analogues

including number of convulsions, general severity (tremors, prostration, dyspnea), duration, and mortality. Both calcium gluconate and sodium gluconate reduced mortality but not severity. Gluconic acid increased the survival time and reduced mortality but did not reduce convulsions or severity. Calcium chloride reduced convulsions but not mortality or tremors. The lifesaving capacity of calcium gluconate at a rate of 40 mg/kg was confirmed by Judah (1949), even though he found normal blood calcium values in most poisoned but unmedicated animals. One animal showed a high calcium value, and Cameron and Burgess (1945) reported a similar result. Calcium has, then, an antidotal action against DDT in intact animals of several species. The hypothesis has been advanced that certain neurotoxins, including DDT, act by delaying the restoration of calcium ions to a surface complex following breaking of the chelate linkage of calcium ions to surface polar groups by an initial exciting impulse (Gordon and Welsh, 1948). This action of the neurotoxin is conceived as depending largely on its physical rather than on its chemical properties. The hypothesis is still helpful in explaining the fact that a wide variety of chemically unrelated compounds produce repetitive responses in excitable tissue and also the fact that many compounds that show a high toxicity for arthropods and mammals are fat-soluble and chemically relatively inert. On the other hand, calcium may help to offset the effects of DDT on calcium-dependent ATPases, especially in the neuronal axons (see Section 93.2.3.7). Having observed the effect of DDT on the metabolism of glucose and glycogen, Lauger and colleagues (Lauger et al., 1945a,b) investigated the use of glucose as an antidote. All of the 10 dogs given 2000 mg of DDT per kilogram

1993

of body weight orally in the form of an oil solution died within 8–24 h. Five of the 10 dogs treated with one or more 20-ml doses of 20% glucose survived the same dosage of DDT. Koster found that glucose given before or after an LD33 dosage reduced convulsions and mortality and, when given before the poison, reduced tremors, prostration, and dyspnea in cats, but was ineffective against an LD95 dosage except to increase the time of survival (Koster, 1947).

93.2.4  Toxicity to Humans 93.2.4.1  Experimental Oral Exposure Table 93.4 summarizes the effects of one or a few oral doses of DDT that are known cases but not of recent history. The results are consistent with those accidents reported in which it was possible to estimate accurately the amount ingested (Garrett, 1947; Hsieh, 1954). It may be concluded that a single dose of 10 mg/kg produces illness in some but not all subjects even though no vomiting occurs. Smaller doses generally produce no illness. Persons who were made sick by 10 mg/kg did not show convulsions, but convulsions have occurred in accidents when the dosage level was 16 mg/kg or greater (Hsieh, 1954). Rarely, a dosage as high as 20 mg/kg has been taken without apparent effect (MacCormack, 1945). Dosages at least as high as 285 mg/kg have been taken accidentally without fatal result (Garrett, 1947). However, large doses lead to prompt vomiting so the amount actually retained cannot be determined accurately. In acute poisoning of volunteers, a slight decrease in hemoglobin and a moderate leukocytosis without any

Table 93.4  Summary of the Effects of One or a Few Oral Doses of DDT on Volunteersa Dose (mg) and formulation

Result

250  9, Suspension

No effect

1500, Butter solution

No effect, but mice killed 6 and 12 h after dose

500, Oil solution

No clinical effect

770, Oil solution

No clinical effect; DDA measured in urine

250, Suspension

None except slight sensitivity of mouth

250, Oil solution

Variable hyperesthia of mouth

500, Oil solution

Variable hyperesthia of mouth

750, Oil solution

Disturbance of sensitivity of lower part of face; uncertain gait; peak reaction after 6 h with malaise, cold moist skin, and hypersensitivity to contact; reflexes normal

1000, Oil solution

Same as above; no joint pains, fatigue, fear or difficulty in seeing or hearing

1500, Oil solution

Prickling of tongue and around mouth and nose 2.5 h after dose; disturbance of equilibrium; dizziness; confusion; tremor of extremities; peak reaction (10 h after ingestion) characterized by great malaise, headache, and fatigue; delayed vomiting; almost complete recovery in 24 h

a

See for original data Domenjoz, 1944; MacCormack, 1945; Neal et al., 1946; Velbinger, 1947a,b.

1994

constant deviation in the differential white count have been observed (Velbinger, 1947a,b). These findings are considered secondary to the neurological effects. In the course of tests with volunteers, dilute colloidal aqueous suspensions of DDT are apparently odorless and tasteless (Domenjoz, 1944; Hoffman and Lendle, 1948). Saturated alcoholic solutions of DDT have a weak aromatic taste, or rather odor. Some people find these solutions slightly anesthetic to the tongue (Hoffman and Lendle, 1948). The taste of DDT in vegetable oil is so slight that many persons cannot identify capsules containing 0, 3.5, and 35 mg of DDT when they are presented separately but can arrange them in proper order when one of each is available for comparison. The possible clinical effects of many repeated doses of DDT were first explored by Fennah (1945). Because of his interest in predicting the results of indiscriminate use, he expressed the exposures in terms of environmental levels rather than in dosage units. The exposures were clearly higher than those ordinarily encountered. In one test, lasting a total of 11.5 months, Fennah daily inhaled 100 mg of pure DDT and drank water dusted at the rate of 3240 mg/m2. Much of the inhaled dust must have been deposited in the upper respiratory tract and swallowed. Later, for 1 month, Fennah ate food all of which had been sprayed at the rate of 2160 mg/m2 after it had been served. No ill effect of any kind was observed. However, these days we would examine more closely possible subtle chronic effects. Some studies of DDT in volunteers have been designed to explore the details of storage and excretion of the compounds in humans and to search for possible effects of doses considered to be safe. In the first of these studies, men were given 0, 3.5, and 35 mg/person/day (Cueto et al., 1956). These administered dosages plus DDT measured in the men’s food resulted in dosage levels of 0.0021–0.0034, 0.038–0.063, and 0.36–0.61 mg/kg/day, respectively, the exact value depending on the weight of each individual. Six volunteers received the highest dosage of technical DDT for 12 months, and three received it for 18 months. A smaller number of men ingested the lower dosage of technical DDT or one of the dosages of p,p-DDT for 12–18 months. No volunteer complained of any symptom or showed any signs of illness that did not have an easily recognizable cause clearly unrelated to the exposure of DDT. At intervals, the men were given a systems review, physical examination, and a variety of laboratory tests. Particular attention was given to the neurological examination and liver function tests, because the major effects of DDT in animals involve the nervous system and the liver. The same result was obtained in a second study in which the same dosages were given for 21 months and the volunteers were observed for a minimum of 27 additional months (Hayes et al., 1971). In the first study, the storage of DDT was proportional to dosage, but there was a then unexplained difference in the storage of the p,p-isomer

Hayes’ Handbook of Pesticide Toxicology

and of technical DDT. Following dosing for 12 months, the pure material was stored in fat at an average concentration of 340 ppm, but the technical material was stored at an average of only 234 ppm. The difference was statistically significant for the 3.5-mg/person/day dosages given for 3–6 and for 7–18 months. The difference was significant for the 35 mg/person/day doses after 7–18 months of dosing but not after only 3–6 months. Men who ate p,p-DDT showed a definite increase in the absolute amount of DDE stored. After 6 months at a dosage of 35 mg/person/day, eight men showed an average DDE fat storage of 32.6  7.0 ppm as compared to 12.3  1.5 ppm for the same individuals upon entering the investigation. There was a further increase of DDE storage as exposure progressed. However, DDT was stored in so much greater concentration that the relative storage of DDE decreased sharply. Thus, after 6 months at a dosage of 35 mg/person/day, eight men stored only 14% of their total DDT-derived material in the form of DDE as compared to 65% for the same persons at the beginning of the investigation. The storage of DDE by men who ate technical DDT presented a different picture. Until 18 months after exposure, there was no clear evidence that these men stored any more DDE after exposure than they did before. However, at 18 months the only three samples available showed DDE concentrations ranging from 28 to 85 ppm, all substantially above general population levels. Thus, both the total amount stored and the rate at which DDT converted to DDE served to distinguish the metabolism of p,p-DDT and technical DDT in humans (Cueto et al., 1956). This was true even though later study showed that the concentration of DDE in serum increased immediately in persons ingesting technical DDT at rates of 10 and 20 mg/person/day. Of course, daily values were subject to considerable variation, but the upward slopes of the graphs recording the results were apparent in 60 days or less and apparently the graphs were straight throughout the 50-month feeding period. Under the same conditions, the level of DDT in serum increased within 1 day and continued to increase in a curvilinear fashion for 5 months (Apple et al., 1970). A similar rapid increase reaching its maximum in 30 h after a single exposure has been observed in workers, as has the more rapid excretion of o,p-DDT (Apple et al., 1970; Morgan and Roan, 1972). In a second study, in which the volunteers received 0, 3.5, and 35 mg/person/day, the storage of DDT was again proportional to dosage with the real but very gradual accumulation of DDE (Hayes, 1982; Smith, 1991). A steady state of storage was approached later in the second study (18.8–21.5 months) than in the earlier one (about 12 months). The second study was superior in that more men were observed for a longer period but inferior in that dosing was less regular. Because of the latter difficulty, it seems impossible to decide whether 12 months or 21.5 months is a more valid estimate of the time necessary for people to approach a steady state of storage

Chapter | 93  Toxicology of DDT and Some Analogues

when intake is uninterrupted and unvarying in amount. It is interesting that the storage levels eventually reached at the same dosage in the two studies were statistically indistinguishable in most instances. In the one instance in which a statistical difference existed, the greater storage by men in the second study may have been explained by the fact that some of them inadvertently received higher doses than intended. DDT was lost slowly from storage in fat after dosing was stopped. The concentration remaining following 25.5 months of recovery was from 32 to 35% of the maximum stored for those who had received 35 mg/person/ day but was 66% for those who had received only 3.5 mg/ person/day, indicating slower loss at lower storage levels. Hayes et al. (1971) and Morgan and Roan (1971) fed volunteers not only technical DDT but also p,p-DDD. They found that DDE is stored more tenaciously than the other compounds in humans, the order being p,p-DDE > p,pDDT > o,p-DDT  p,p-DDD. The slow metabolism of DDT to DDE was confirmed. It was noted that p,p-DDT is lost from storage in adipose tissue much more slowly in humans than in the monkey, dog, or rat. Less than 18% of p,p-DDE is carried in human erythrocytes. In plasma of ordinary fat content, less than 1% of all DDT-related compounds is carried by the chylomicrons. p,pDDT and p,p-DDE are found mainly in the triglyceride-rich, low-density, and very low-density lipoproteins. Following continuous electrophoresis, these compounds are found mainly in association with plasma albumin and -globulins (Morgan and Roan, 1972). DDA is the main urinary metabolite of DDT. In humans, it was found first in a volunteer by Neal et al. (1946) who reported that, following ingestion of 770 mg of p,p-DDT, excretion rose sharply to 4.0 mg/day during the second 24-h period, decreased rapidly on the 3rd and 4th days, decreased gradually there­ after, but was still above baseline on day 14. Later studies in volunteers who received smaller but repeated doses confirmed the very rapid rise in excretion of DDA. Hayes et al. (1971) and Roan et al. (1971) showed that a steady state of excretion was reached after about 6–8 months. During a 56-week period of continued dosing after equilibrium was fully established, the concentration of DDA associated with technical DDT at the rate of 35 mg/person/ day varied from 0.18 to 9.21 ppm and averaged 2.98 ppm; corresponding values for p,p-DDT were 0.40–6.27 ppm with a mean of 1.88 ppm. Thus technical DDT, as compared to p,p-DDT, was excreted more effectively and stored less. During the latter half of the dosing period, in the two groups receiving recrystallized and technical DDT at the rate of 35 mg/person/day, it was possible to account for an average of 13 and 16%, respectively, of the daily dose in terms of urinary DDA. The excretion of DDA was relatively constant in each individual, but marked differences were observed between men receiving the same dose (Hayes et al., 1971; Smith, 1991).

1995

93.2.4.2  Experimental Dermal Exposure Depending on dosage, oral administration of DDT to volunteers has produced either no illness or brief poisoning entirely similar to that seen in experimental animals. The oral dosage necessary to produce any clinical effect was almost always 10 mg/kg or more. It is a strange coincidence that, in two studies involving only three subjects in all, experimental dermal exposure to DDT was followed by fatigue, aching of the limbs, anxiety or irritability, and other subjective complaints. Recovery was delayed 1 month or more (Case, 1945; Wigglesworth, 1945). In neither study was there an independent control. Although the dosage was unmeasured, the amounts of DDT absorbed must have been much smaller than those involved in the oral studies. One of the studies involved self-experimentation by one man. A somewhat more severe test on six volunteers produced no toxic or irritant effect at all (Dangerfield, 1946). In view of all other experiments and extensive practical experience, it is probable that the illnesses reported by Wigglesworth and Case were unrelated to DDT. With the exceptions just mentioned, dermal exposure to DDT has been associated with no illness and usually no irritation (Cameron and Burgess, 1945; Chin and T’Ant, 1946; Dangerfield, 1946; Domenjoz, 1944; Draize et al., 1944; Fennah, 1945; Haag et al., 1948). In fact, Hoffman and Lendle (1948) reported that even subcutaneous injection of colloidal suspensions of DDT in saline in concentration up to 30 ppm caused no irritation. Zein-el-Dine (1946) reported that DDT-impregnated clothing caused a slight, transient dermatitis, but the method of impregnation was not stated and the absence of solvent was not guaranteed. Other more thorough studies of DDT-impregnated clothing have found it nonirritating (Cameron and Burgess, 1945; Domenjoz, 1944). Chin and T’Ant applied small pads impregnated with different formulations of DDT to the inner surface of the forearm of 32 volunteers whose cutaneous sensation had previously been measured for a period of 5 weeks (Chin and T’Ant, 1946). Pads impregnated with all the elements of the formulation except DDT were applied to the corresponding position on the other arm as a control. Powdered DDT and 5% solutions of DDT showed little effect. Ten percent and 20% solutions in olive oil and petroleum showed no remarkable effect on sensation of pain, cold, or heat but reduced tactile sensation in most cases so that the minimal pressure that could arouse this sensation was 1–2.5 gm/cm2 higher than in the control.

93.2.4.3  Experimental Respiratory Exposure Neal et al. reported almost continuous daily exposure to aerosols sufficient to leave a white deposit of DDT on the nasal vibrissae of the volunteers (Neal et al., 1944). This exposure produced moderate irritation of the nose, throat,

1996

and eyes. Except for this irritation during exposure, there were no symptoms, and laboratory tests and physical examination, including neurological evaluation, failed to reveal any significant changes. The studies by Fennah that involved both respiratory and oral exposure, produced no detectable ill effect, as discussed above (Fennah, 1945). Stammers and Whitfield reported tests in which volunteers were exposed to DDT dispersed into the air either by volatilizing units or by continuously or intermittently operated aerosol dispensers (Stammers and Whitfield, 1947). In some instances, a slight odor and some dryness of the throat were noticed, but otherwise the results were negative.

93.2.4.4  Therapeutic Use The early use of DDT for treating human body lice, head lice, and scabies was reviewed by Simmons (1959). Obviously, these uses offered a possibility of dermal absorption, but such absorption of dry DDT is very limited. Persons who had DDT blown into their clothing as they wore it must have inhaled some of the compound, and this was especially true of persons who used hand or power equipment to apply the dust to hundreds of people per day in mass delousing stations set up to control typhus (West and Campbell, 1946). However, the dosages absorbed cannot have been so large as in some instances in which DDT has been administered by mouth. Even smaller absorbed dosages for the general population were involved in the use of DDT for the control of other vector-borne diseases, especially malaria. These uses of DDT made tremendous contributions to human health through control of the vectors of typhus, malaria, plague, and several lesser diseases (Coulston, 1985a; Hayes and Laws, 1991; Spindler, 1983). DDT has been used on an experimental basis at oral dosage rates varying from 0.3 to 3 mg/kg/day for periods up to 7 months in an attempt to decrease serum bilirubin levels in selected patients with jaundice (Thompson et al., 1969). No side effects were observed. No improvement was noted in patients with jaundice based on cirrhosis who had no demonstrated liver enzyme deficiency. However, in a patient with familial, nonhemolytic, unconjugated jaundice based on a deficiency of glucuronyltransferase, treatment with DDT rapidly reduced the plasma bilirubin level to the normal range and relieved the patient of nausea and malaise from which he had suffered intermittently. The liver function tests as well as other laboratory findings remained normal. The improvement was maintained during the 6 months when DDT was administered and had persisted for 7 additional months at the time the report was written. In this case, a dosage of 1.5 mg/kg/day produced a steady rise in plasma levels of p,p-DDT from an initial level of 0.005 ppm to a maximum of 1.33 ppm at the end of treatment. At this time, the concentration in body fat was 203 ppm. Plasma levels fell slowly after dosing was stopped (Thompson et al., 1969). The highest daily intake

Hayes’ Handbook of Pesticide Toxicology

in this series was six times greater than the highest level administered in earlier studies of volunteers and about 7500 times greater than the DDT intake of the general population. These days the difference would be even greater. The highest value for p,p-DDT in serum observed in the entire series was 1.330 ppm, compared to 0.996 ppm, the highest value reported by Laws et al. (1967) for formulation-plant workers. A lesser induction of the microsomal enzymes has been observed in workers also (Kolmodin et al., 1969; Poland et al., 1970). Rappolt used a single dose of 5000 mg of DDT to promote the metabolism of phenobarbital, of which his three patients had taken an overdose (Rappolt, 1970). The treatment appeared useful. Neither Rappolt nor Thompson encountered any side effects of DDT (Rappolt, 1970; Thompson et al., 1969). However, in addition to whatever action it may have had in promoting the metabolism of phenobarbital, the DDT administered by Rappolt must have acted largely as a pharmaceutical antidote for the barbiturate. The largest dose previously administered intentionally was 1500 mg, which caused moderate poisoning in a volunteer, who, of course, had received no barbiturate (Table 93.4).

93.2.4.5  Accidental and Intentional Poisoning The earliest symptom of poisoning by DDT is hyper­ esthesia of the mouth and lower part of the face (Eskenasy, 1972; Hayes, 1982). This is followed by paresthesia of the same area and of the tongue and then by dizziness, an objective disturbance of equilibrium, paresthesia and tremor of the extremities, confusion, malaise, headache, fatigue, and delayed vomiting. The vomiting is probably of central origin and not due to local irritation. Convulsions occur only in severe poisoning. Onset may be as soon as 30 min after ingestion of a large dose or as late as 6 h after smaller but still toxic doses. Recovery from mild poisoning usually is essentially complete in 24 h, but recovery from severe poisoning requires several days. In two instances, there was some residual weakness and ataxia of the hands 5 weeks after ingestion. Involvement of the liver has been mentioned in only a small proportion of cases of accidental poisoning by DDT. In three men who ate pancakes made with DDT and who ingested 5000–6000 mg each, slight jaundice appeared after 4–5 days and lasted 3–4 days (Naevested, 1947). Hepatic involvement and convulsions were reported in an unsuccessful attempt at suicide by ingesting DDT and lindane (Eskenasy, 1972). Cases of individual and suicidal poisoning in which effects were clearly caused by DDT ingestion are summarized in Table 93.5. The signs and symptoms of poisoning were entirely consistent with those observed in volunteers, except that the spectrum of effects was broader because some of the accidental and suicidal doses were very high. A few persons apparently have been killed by

Chapter | 93  Toxicology of DDT and Some Analogues

1997

Table 93.5  Summary of the Effects of the Accidental or Suicidal Ingestion of DDT Individual dose (mg), formulation, and number of persons

Findings

300–4500, in food, one man

Onset in 1 h; vomiting; restlessness; headache; heart weak and slow; recovery next day (Morgan and Hickenbottom, 1979)

Unknown dose, in tarts, 25 men

Onset in 2–2.5 h; all weak and giddy; four vomited; two hospitalized; one confused, incoordinated, weak; one with palpitations and numbness of hands; recovery in 24–48 h (Mackeras and West, 1946)

5000–6000, in pancakes, three men

Onset 2–3 h; throbbing headache; dizziness; incoordination; paresthesias of extremities; urge to defecate; wide nonreacting pupils; reduced vision; dysarthria; facial weakness; tremor; ataxic gait; reduced sensitivity to touch; reduced reflexes; positive Romberg; slightly low blood pressure and persistent irregular heart action; partial recovery in 2–3 days, but slight jaundice appeared 4–5 days after ingestion and lasted 3–4 days; all normal 19 days after poisoning except irregular heart action in one (Naevested, 1947)

2000, in pancakes, two men

No illness (Naevested, 1947)

Up to 20,000, in bread, 28 men

Onset in 30–60 min in those most severely affected, men first seen 2–3 h after ingestion; in spite of severe early vomiting that reduced the effective dose, severity of illness and especially intensity of numbness and paralysis of extremities proportional to amount of DDT ingested; all but eight men recovered in 48 h; five others fully recovered in 2 weeks, but three men still had some weakness and ataxia of their hands 5 weeks after ingestion (Garrett, 1947)

Unknown dose, in flour, about 100 women

Onset about 3.5 h after ingestion; total of about 85 cases of which 37 were hospitalized; symptoms mild and similar to those in earlier outbreaks except gastrointestinal disturbance in most severe cases included abdominal pain and diarrhea as well as nausea; most fully recovered in 24 h (Jude and Girard, 1949)

Unknown dose, 14 cases

Symptoms in established cases similar to those reported earlier (Francone et al., 1952)

286–1716, in meatballs, eight cases, 11 exposed

With the exception of one man who was already sick when he received a dosage of 6 mg/kg, poisoning did not occur at dosages of 5.1–10.3 mg/kg. Ingestion of 16.3– 120.5 mg/kg produced excessive perspiration, nausea, vomiting, convulsions, headache, increased salivation, tremors, tachycardia, and cyanosis of the lips. Onset varied from 2 to 6 h, depending on dosage. Recovery required as much as 2 days (Hsieh, 1954)

Unknown dose, one case

Death 13 h after suicidal ingestion (Committee on Pesticides, 1951 quoted by Hayes, 1982)

Unknown dose, 22 unrelated cases

22 Separate cases, including 15 attempted suicides; some complicated by solvents; three deaths (Committee on Pesticides, 1951 quoted by Hayes, 1982)

uncomplicated DDT poisoning, but none of these cases was reported in detailed literature. Often death attributed to DDT has been caused by the ingestion of solutions of DDT, but in most of these instances the signs and symptoms were predominantly or exclusively those of poisoning by the solvent (Hayes, 1959). Sometimes findings have been ambiguous. For example, the recurrent convulsions in one case reported (Cunningham and Hill, 1952), though more characteristic of poisoning by one of the cyclodienes, was certainly not typical of solvent poisoning. A 2-yearold child drank an unknown quantity of fly spray of which 5% was DDT, but the nature of the other active ingredients or the solvent was unknown. About 1 h after consumption, the child became unconscious and had a generalized, sustained convulsion. Convulsions were present when the

child was hospitalized 2 h after taking the poison, but the fits were controlled by barbiturates and other sedatives. Convulsions reoccurred on day 4 and again on day 21 but were stopped each time following renewal of treatment. On day 12, it was noted that the patient was deaf. Hearing began to improve about day 24 and was normal, as were other neurological and psychic findings, when the patient was seen about 2.5 months after the accident. Clinical effects of one toxicant may be modified by combining it with another. For example, one would not expect prolonged illness from DDT at a rate of 27 mg/kg. However, when DDT and lindane were ingested in a suicidal attempt at dosages thought to be 27 and 18 mg/kg respectively, clinical remission of convulsions and of liver involvement was delayed until day 20, and the EEG did not

1998

return to normal until day 39 (Eskenasy, 1972). What little is known about the effect of DDT on the human heart fails to show whether cardiac arrhythmia might be a possible cause of death in acute poisoning, as is true in some species of laboratory animals. Palpitations, tachycardia, and “irregular heart action” have been noted in some but not all cases of acute poisoning (Hsieh, 1954; Mackeras and West, 1946; Naevested, 1947). There do not seem to be any accidents or suicides involving respiratory or dermal exposure leading to recognized signs and symptoms of DDT poisoning. This is true even though sufficient respiratory exposure to aerosols or sufficient dermal exposure to solutions can cause poisoning in animals, and the difference is certainly one of dosage.

93.2.4.6  Use and Environmental Exposure Despite its bad press and concern over its environmental impact, the safety record for humans in the use of DDT is phenomenally good considering the huge quantities distributed (Coulston, 1985a; Smith, 2000). It has been used for mass delousing in such a way that the bodies and inner clothing of thousands of people of all ages and states of health were liberally dusted with the compound (West and Campbell, 1946). By necessity, the applicators worked in a cloud of the material. Other applicators have sprayed the interior of hundreds of millions of homes in tropical and subtropical countries under conditions that Wolfe et al. (1959) showed involved extensive dermal and respiratory exposure. A smaller number of people have made or formulated DDT for many years. Extensive experience and numerous medical studies of groups of workers at a time when vast quantities were manufactured have been reviewed (Hayes, 1959). Dermatitis was commonly observed among workers who used DDT solutions. The rashes were clearly due to the solvent, especially kerosene. As often happens with rashes caused by petroleum distillates, they were most severe in people when they first started work and cleared in a few days unless contamination was exceptionally severe. A smaller number of workers experienced mild narcotic effects (vertigo and nausea) from solvents when working in confined spaces. Some persons suffered temporary irritability, fatigue, and other ill-defined symptoms after exposure in the dusty atmosphere of a delousing station, but the relation of these atypical findings to DDT was not clear (Gil and Miron, 1949). With these exceptions due largely to solvents, no illnesses clearly attributable to the formulations, much less to DDT, were revealed by the early studies. Mild to moderate poisoning by DDT itself may have occurred among a group of factory workers exposed to air concentrations of 5–4200 mg/m3, but no measurements were made of DDT in blood, fat, or urine. The workers complained of paresthesia of the extremities, headache, dizziness, and some other difficulties less clearly linked

Hayes’ Handbook of Pesticide Toxicology

to DDT (Aleksieva et al., 1959). Even higher concentrations in air have been associated with tremor of the tongue and hands as well as with numerous subjective findings (Burkatzkaya et al., 1961). Ortlee carried out clinical and laboratory examinations of 40 workers, all of whom were exposed to a number of other pesticides (Ortelee, 1958). They had been employed at this work, with heavy exposure, for 0.4–6.5 years with slightly less exposure for as much as 8 years. Exposure was so intense that during working hours many of the men were coated with a heavy layer of concentrated DDT dust. By comparing their excretion of DDA with that of volunteers given known doses of DDT, it was possible to estimate that the average absorbed dosages of three groups of the workers with different degrees of occupational exposure were 14, 30, and 42 mg/person/day. With the exception of the excretion of DDA and the occurrence of a few cases of minor irritation of the skin and eyes, no correlation was found between any abnormality and exposure to the insecticide. Since very large doses of DDT injure the nervous system and liver of experimental animals, special attention was given to a complete neurological examination and to laboratory tests for liver function. Although a few abnormalities were revealed, none related to DDT was detected. Laws studied 35 men employed from 11 to 19 years in a plant that had produced DDT continuously and exclusively since 1947 and, at the time of the study, produced 2722 metric tons per month (Laws et al., 1967). Findings from medical history, physical examinations, routine clinical laboratory tests, and chest X-ray films did not reveal any ill effects attributable to exposure to DDT. No case of cancer or blood dyscrasia was found among the 35 heavily exposed workers in a DDT factory, nor did the medical records of 63 men who had worked there for more than 5 years reveal these diseases (see Section 93.2.4.7). Two men were employed who had a history of successfully treated cancer before they came to work, but no employee had contracted cancer during the 19 years the plant had operated; during this period, the workforce varied from 110 to 135. A study of liver function of the heavily exposed men is discussed near the end of this section. Measurement of storage offered direct evidence of the men’s heavy exposure. The overall range of storage of the sum of isomers and metabolites of DDT in the men’s fat was 38–647 ppm, compared to an average of 8 ppm for the general population. Based on their storage of DDT in fat and excretion of DDA in urine, it was estimated that the average daily intake of DDT by the 20 men with high occupational exposure was 17.5–18 mg/person/day, compared to an average of 0.028 mg/person/day then found for the general population. There was significant correlation between the concentrations of total DDT-related material in the fat and serum of the workers. The concentration in fat averaged 338 times greater than that in serum, a factor about three times greater than that for people without occupational

Chapter | 93  Toxicology of DDT and Some Analogues

exposure. Compared to people in the general population, workers were found to store a smaller proportion of DDTrelated material in the form of DDE; the difference was shown to be related chiefly to intensity rather than duration of exposure. DDE is relatively much less important and DDA much more important as excretory products in occupationally exposed men than in men of the general population. After Laws et al. (1967) had completed their study, it was found that the 36 most heavily exposed workers involved had fathered 58 children before they began working at the DDT factory and 93 children afterward (Wilcox, 1967, quoted by Hayes, 1982; Smith, 1991, 2001). Laws and coworkers made a detailed study of the liver function of 31 men who had made and formulated DDT and who had been the subjects of the earlier study already discussed (Laws et al., 1973). Judging from their excretion and storage, the men’s exposure was equivalent to an oral intake of DDT at rates ranging from 3.6 to 18 mg/man/day for periods ranging from 16 to 25 years and averaging 21 years. All tests were in the normal range for total protein, albumin, total bilirubin, thymol turbidity, and retention of sufobromophthalein sodium. One man had mild elevation of alkaline phosphatase and serum glutamic-pyruvic transaminase (SGPT). Another man had elevated alkaline phosphatase concentration of 14 units, while a third man had an elevated SGPT. Comparison of the residue levels of DDE, o,p-DDT, and p,p-DDT in blood of factory workers exposed to DDT formulations and showing apparent temporary clinical symptoms and those without symptoms showed no significant differences. Levels were approximately 10-fold greater than those of unexposed controls (Chand et al., 1991). By far the largest number of heavily exposed workers whose health has been investigated was those associated with malaria control (WHO, 1973). In Brazil, periodic clinical examinations were made of 202 sprayers exposed to DDT for 6 or more years, 77 sprayers exposed for 13 years ending in 1959, and 406 controls. During a 3-year period, a survey of illnesses requiring medical care during the 6 months preceding each periodic medical examination failed to demonstrate any difference between exposed and control groups. A small number of analyses indicated that the concentration of DDT in the blood of sprayers was about three times higher than that of controls. In India, the blood levels of 144 sprayers were 7.5–15 times greater than those in controls and were at least as high as those reported for workers who had made and formulated DDT elsewhere (Misra et al., 1984). When the sprayers were examined, no differences from controls were found except that knee reflexes were brisker, slight tremor was more often present, and a Romberg test was more poorly performed by the sprayers. The positive results led to the selection of 20 men for reexamination by a neurologist, who concluded that the differences found initially were not real or that the tests had returned to normal within the few months between the two examinations. In any event, the signs were not dosage-related, since they showed no correlation

1999

with serum levels of DDT. Subsequently cognitive functions of Indian DDT sprayers were tested and DDT levels were 8.5 times higher than those in controls and visuomotor functions were significantly depressed. In an association study, the GSTP1 val-105 allele in children has been linked with higher risk of adverse cognitive function of prenatal DDT exposure (Morales et al., 2008). Perhaps in contrast to the above studies, the levels of DDT and its metabolites in the sera from 23 applicators in malaria control in Natal were significantly higher than in the population protected by the spraying (Bouwman et al., 1991b). Although serum GGT was not statistically different from controls, the mean in applicators was greater than the maximum laboratory mean level and ALT values were significantly greater in the applicators although not deemed clinically significant. Members of households which had been sprayed internally with DDT had significantly greater levels in their serum than people from nonsprayed households (sum of DDT and metabolites was 140.0 g/l compared to 6.4 g/gl). Although GGT levels were greater in the high DDT group this seemed to be associated with alcohol consumption (Bouwman et al., 1991a). The induction by DDT level of microsomal enzymes of human liver was demonstrated first in workers (see Section 93.2.4.4) and DDT may be more important than DDE in this regard, as indicated by the fact that Poland et al. (1970) observed induction in men with average serum levels of 0.573 and 0.506 ppm for DDT and DDE, respectively, while Morgan and Roan found no induction in men with corresponding values of 0.052 and 0.222 ppm (Morgan and Roan, 1974b). As noted in Section 93.2.4.4, DDT has been used successfully to induce microsomal enzymes in order to promote metabolism of bilirubin in a case of congenital defect and to promote metabolism of phenobarbital in a case of overdose. DDT promotes its own metabolism in some species of laboratory animals. That the same is true in humans is indicated by the fact that storage of DDT is relatively less at higher dosages. However, the metabolism and subsequent excretion of DDT can be promoted even more by phenobarbital and especially diphenylhydantoin (see Section 93.2.3.5) and by some other drugs (McQueen et al., 1972). Establishment of a reduced equilibrium appeared to require about 2 months. Within this period, the regression of the level of DDT plus DDE on duration of treatment with diphenylhydantoin was highly significant. In addition to the studies already mentioned regarding workers with extensive storage and/or excretion of DDT as a result of truly heavy exposure to DDT, studies also have been made of a larger number of workers with lesser storage and/or excretion following lesser exposure to DDT but greater exposure to other insecticides. Continuing, meticulous study discussed by Hayes (1982) and in the previous edition of this chapter (Smith, 2001), as well as the work of other investigators (Morgan and Lin, 1978; Ouw and

2000

Shandar, 1974), failed to reveal effects of clinical significance among workers with prolonged, moderate exposure to a wide variety of pesticides. A review of results for 2620 persons exposed to pesticides and 1049 persons not occupationally exposed (Morgan and Lin, 1978) found that, apart from serum pesticide concentrations, the only significant and consistent change associated with occupational exposure was a depression of serum bilirubin. This presumably was a reflection of a slight induction of liver microsomal enzymes. In addition, there was a tendency for serum alkaline phosphatase to increase with increasing concentrations of DDT plus DDE in the serum, but the differences were small in all instances and statistically significant for SGPT, serum glutamic oxaloacetic transaminase (SGOT), and lactate dehydrogenase (LDH) only. Wong could find no significant overall cause specific mortality excess among men potentially exposed at work to DDT from 1935 to 1976 (Wong et al., 1984). Similarly, a population of 499 persons living downstream from a defunct DDT-manufacturing plant showed no DDT-specific illnesses or ill health despite total DDT serum levels three times the national mean (Kreiss et al., 1981). There was, however, a possible association between serum DDT and serum cholesterol, triglyceride, and -glutamyl transpeptidase levels. A positive linear correlation has been reported for the concentrations of vitamin A and of DDT-related compounds in the serum of men with at least 5 years of occupational exposure to DDT. However, the workers’ DDT levels were little higher than those of persons in the general population (see Table 15.11 in Smith, 1991) and their vitamin A levels were within normal limits (Keil et al., 1972a). One epidemiology study reports an association of lower T3 levels in blood of 4-year-old children with DDT but also with other chemicals (Alvarez-Pedrerol et al., 2008).

93.2.4.7  Evidence for Causing Cancer Evidence regarding mutagenic activity of DDT and its significance is humans is uncertain. Comparing samples collected in winter and during the peak season of pesticide application, a slight increase in chromatid breaks was reported in the cultured lymphocytes of workers exposed to a wide variety of insecticides said to include DDT (Yoder et al., 1973). A somewhat larger increase was reported for men exposed mainly to herbicides (You et al., 1998). The paper failed to explain why exposure to DDT was claimed at a time when its use was banned. In another study, lymphocytes cultured from workers with an average DDT plasma level of 0.999 ppm showed significantly more chromosomal and chromatid aberrations than did cells cultured from controls with an average plasma level of 0.275 ppm (Rabello et al., 1975). The difference was not significant in other comparisons in which the average plasma levels were 1.030 vs. 0.380 ppm and 0.240 vs. 0.030 ppm, respectively. Examination of all of the data presented by the authors suggests that a simple dosage–effect relationship was

Hayes’ Handbook of Pesticide Toxicology

present, with a detectable effect starting somewhere between 0.2 and 0.4 ppm and increasing at levels higher than 0.4 ppm. Some chromosomal aberrations have also been observed with human lymphocyte cultures by Preston et al. (1981), but DDT did not cause unscheduled DNA synthesis in SV40-transformed human cells (Ahmed et al., 1977). In Denmark, Unger found significantly higher levels of DDE in adipose tissue from terminal cancer patients than in tissue from patients who died from other causes (Unger and Olsen, 1980). In the United States, DDT and DDE levels were measured in 919 subjects in 1974 and 1975 (Austin et al., 1989). After 10 years there was no correlation between these levels and overall mortality or cancer mortality except a slight correlation with respiratory cancer death. Of course, increased storage often correlates with emaciation of whatever cause. A study has been made of 1043 deaths that occurred between 1956 and 1992 among men who used DDT in an antimalarial campaign in Sardinia in the late 1940s (Cocco et al., 1997). Workers had a significant increased risk for liver and biliary tract cancers and multiple myeloma. However, nonexposed workers also showed elevated incidences of cancer. An increased incidence of liver cancer associated with previously higher plasma levels of DDE in U.S. citizens (in 1968) has been noted (Cocco et al., 2000). There have been reports that pancreatic cancer might be associated with exposure to DDT and ethylan. In a nested case–control mortality study among 5586 workers at a chemical plant, interviews with next of kin and coworkers and examination of work records showed that DDT exposure appeared to be associated with pancreatic cancer as identified by pathology, clinical surgical, autopsy records, or death certificates (Garabrant et al., 1992). Among subjects whose mean exposure to DDT was 47 months, the risk was 7.4 times that among subjects with no exposure. DDD and ethylan exposures were also correlated. The validity of the classification of pancreatic cancer was subsequently questioned (Malats et al., 1993). It was proposed that despite the difficulties, it might strengthen the findings if comparisons were made between histologically confirmed cases. Interestingly, this indeed turned out be true (Garabrant et al., 1993). For those cases in which pancreatic cancer classification was based only on death certificates, no association with exposure to DDT or its analogues was found. This theme was developed further by looking at 66 residents of Michigan (39–70 years of age) who had been diagnosed histologically as having pancreatic cancer (Fryzek et al., 1997). Exposure was assessed by self-reporting questionnaires. In this cohort from the general population, a significant increased risk for exposure to ethylan was observed. Cases were 10.7 times more likely to report exposure to ethylan compared to controls. Nonsignificantly, increased odds ratios were for exposure to DDT and chloropropylate. Examination of a group of cases of exocrine pancreatic cancer has recorded an association of higher serum DDT levels and mutations in the K-ras gene, not found in controls (Porta et al., 1999).

Chapter | 93  Toxicology of DDT and Some Analogues

Present understanding on the mode of action of chemicals causing rodent tumors suggests that DDT per se is unlikely to be a human hepatocarcinogen (Holsapple et al., 2006). Although there is a lot of evidence against DDT causing liver cancer in humans in Western countries (Turusov et al., 2002), there is still the outside possibility of it acting as a promoter of potent carcinogens. Aflatoxin is a well-known human carcinogen in areas of Southeast Asia such as Thailand, where DDT and other chlorinated insecticides have been used more recently. In the past decade there has been continued speculation and controversy as to whether many environmental chemicals, including chlorinated pesticides, can act as so-called endocrine disrupters. As with all environmental exposures, this is an extremely difficult issue to develop to firm conclusions. In some studies, DDT was particularly targeted as being linked with a rising incidence of breast cancer (Wolff, 1995). Much of the evidence has been reviewed in detail (Ahlborg et al., 1995; Calle et al., 2002; Turusov et al., 2002) but is not supportive or is inconclusive. Unger found no relationship between breast fat tissue DDT (and DDE) and the incidence of mammary cancer (Unger et al., 1984). Although in some studies (Mussalo-Rauhamaa et al., 1990) correlation with -hexachlorocyclohexane was found, none was observed with DDT. In contrast, elevated levels of DDT or DDE were reported in cancerous breast tissue fat compared with tissue from benign mammary disease (Falck et al., 1992; Guttes et al., 1998). Studies on the relationship between blood levels of DDT and/or DDE and breast cancer in the United States have been mixed. Some have shown a relationship with either blood p,p-DDE levels and mammary cancer occurrence (Wolff et al., 1993) or p,p-DDE levels and hormone-responsive breast cancer (Dewailly et al., 1994a,b). It has been proposed that increased estrogen receptor level in breast tumors over two decades could be explained by organochlorine exposure (Dewailly et al., 1997). In contrast, a number of studies have found no relationship between blood levels of DDE and risk of breast cancer (Hunter et al., 1997; Krieger et al., 1994; Schecter et al., 1997). The last study is of additional interest in that women in North Vietnam were examined who had generally high levels of DDT or DDE due to exposure from antimalarial use. Many of the studies conducted up to 1994 were examined by meta-analysis and appeared to confirm the lack of association between DDE levels in tissues and breast cancer incidence (Key and Reeves, 1994). However, there are many in vitro and epidemiological investigations since that purport to agree with the hypotheses (e.g., Ardies and Dees, 1998; Shekhar et al., 1997) and others that do not (Iwasaki et al., 2008). Again much of the evidence has been reviewed in detail and is not supportive of a link between DDT exposure and breast cancer. However, an interesting development has been the observation that it may be early in life that exposure to DDT is the most

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significant for breast cancer development (Cohn et al., 2007). Undoubtedly, this debate will continue (Tarone, 2008a,b). Probably there will be a further concern around p,p-DDE acting as an antiandrogen (Kelce et al., 1995, 1998) (see Section 93.2.3.12). This might be associated with apparent risks for testicular germ cell tumors (McGlynn et al., 2008). Rates of non-Hodgkin’s lymphoma (NHL) in the United States and other Westernized countries are usually considered to have increased over the last decades (Fisher and Fisher, 2004). In the United States, it is more common in men than women and more common in whites than blacks. NHL appears to be in fact a group of closely related diseases, probably linked to immune suppression. Some studies have shown high rates of NHL among agricultural workers and people working with pesticides. The problems in determining associations with risk factors continue to have low statistical power because of relatively low numbers, poorly defined exposures often relying on self-reporting of years previously, and the large mixtures of chemicals encountered (Fisher and Fisher, 2004). Although studies have often been positive, it is not possible to exclude confounders such as UV radiation and acquisition of animal pathogens as well as the influence of allergic conditions such as asthma (Lee et al., 2004). Various studies have set out to find whether by environmental exposure of the general population or by exposure in agricultural practices DDT was associated with increased incidences of NHL. However, in a large study of DDE levels in subcutaneous fat from people in 22 U.S. states measured in 1968 and subsequent development of disease from 1975 to 1994, no association with incidence of NHL was found (Cocco et al., 2000). In a nested study of serum DDT and metabol­ ite levels of patients identified from an original 25,802 population in MD sampled in 1974, no correlation with NHL was found either (Rothman et al., 1997). Similarly, no strong consistent evidence was found for association between exposure to DDT and NHL in 993 farmers from midwest states (Baris et al., 1998). In none of the weak associations that been reported for the United States and Italy have clear exposure data been available. It is always impossible to rule out the effects of other pesticides (Cantor et al., 1992; McDuffie et al., 2001; Nanni et al., 1996; Woods et al., 1987). Thus sometimes the apparent significant effects of DDT disappear when exposure or use of other pesticides is taken into account (Baris et al., 1998). In addition, exposure to farm animal pathogens and Epstein-Barr virus might also be confounders (Nanni et al., 1996; Rothman et al., 1997). One consequence, however, is that it may not ever be possible to rule out the action of DDT in a mixture with other pesticides rather than putting the emphasis on DDT acting independently. More recent studies claim that plasma DDE levels in farmers correlate with significant modulation of immune responses (Cooper et al., 2004).

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93.2.4.8  Dosage Response The clinical effects of different dosage levels of DDT in humans are summarized in Tables 93.4 and 93.5. The degree of storage determined by different dosage levels of DDT has been summarized in Figure 7.4 in Levine (1991) and details regarding higher than normal dosage rates are given in Table 15.9 in Smith (1991). A clinically useful degree of induction of microsomal enzymes was obtained with a DDT dosage of 1.5 mg/kg/day for 6 months (see Section 93.2.4.4). As discussed in Section 93.2.4.6, workers who absorbed a dosage of about 0.25 mg/kg/day showed demonstrable but only slight induction (Poland et al., 1970). Workers with less exposure as indicated by lower serum levels of DDT showed no detectable induction.

93.2.4.9  Storage in Fat The highest reported storage of DDT and related compounds remains that of a healthy worker whose fat contained DDT and DDE at concentrations of 648 and 483 ppm, respectively (Cueto et al., 1956). Laws and coworkers reported considerably lower storage values among the most exposed persons in a DDT manufacturing plant (Laws et al., 1967) (see also Table 15.11 of Smith, 1991). An important point is that whereas almost all investigations of workers are said to have been carried out on “heavily exposed” populations (or words to that effect), some of the groups studied had absorbed little more DDT than is absorbed by the general population – especially the general population of some tropical countries. The first evidence that humans metabolize a part of the DDT they absorb to DDE was obtained from the analysis of fat from a worker (Mattson et al., 1953). The accumulation of DDE relative to total DDT-related compounds is best illustrated in humans. Of the total DDT stored in the fat of workers exposed to technical DDT (about 4% DDE) for 11–19 years, only 38% was in the form of DDE, and, of course, some of that DDE came from their diets including meat (Laws et al., 1967). In India, where many people avoid meat but may consume milk, cheese, and eggs, 34–41% of total DDT stored by people without special exposure was DDE (Dale et al., 1965). In the United States, during a time when DDT residues in food were decreasing, the proportion of total DDT in the form of DDE increased from about 60% in 1955 to about 80% in 1970; during the same interval the concentration of total DDT in body fat decreased from about 15 ppm to less than 10 ppm as recorded in Table 7.10 in Hayes (1975). By 1980, DDE constituted 86.7% of total DDT in one population (Kreiss et al., 1981). Thus, a low proportion of DDE indicates a relatively high intake of preformed DDT and relatively few years for metabolism of stored DDT to DDE. A number of factors, especially dosage, age, sex, race, and various disease states have been discussed in connection

Hayes’ Handbook of Pesticide Toxicology

with the storage and excretion of DDT by people but only dosage has been shown to be of practical importance. DDT and related compounds are stored at much lower rates in the general population than in persons with occupational exposure. However, these relatively low levels of storage constitute one of the most important aspects of the measurable effects of pesticides on people (Dale et al., 1963). Briefly, storage of total DDT in the body fat of ordinary people in the United States increased from 5.3 ppm in 1950 to about 15.6 ppm in 1955 and 1956 (see Table 15.12 in Smith, 1991). Thereafter, the levels decreased gradually (Burns, 1974) albeit somewhat irregularly, to about 8 ppm in 1970 and 3 ppm in 1980. In fact, despite concerns about residues of DDT, levels have continued to fall. In annual surveys in the United States based on 898–1920 samples per year, the geometric mean levels for total DDT in adipose tissue on a lipid basis were 7.88, 7.95, 6.88, 5.89, and 5.02 ppm for fiscal years 1970, 1971, 1972, 1973, and 1974, respectively. For each year, the values were higher for older age groups and higher for black than for white people. During fiscal year 1974, the values for persons 0–14, 15–44, and 45 years old or more were 2.15, 4.91, and 6.55 ppm, respectively, for white people and 4.02, 9.18, and 11.91 ppm, respectively, for black people (Kutz et al., 1977). The values would have been somewhat lower if they had been based on wet weight. It has been calculated that if exposure to DDT ceased it would take 10–20 years for DDT to disappear from a person but that DDE would persist throughout the life span (Morgan and Roan, 1974a).

93.2.4.10  Storage in Blood and Other Organs No information seems to be available on blood levels of DDT in persons poisoned by the compound. Concentrations measured in the blood or serum from workers are shown in Table 15.11 of Smith (1991). The highest value for total DDT in serum reported from several countries was 2.2 ppm (with an average of 0.7371 ppm) based on gas chromatography (Laws et al., 1967). Other studies conducted at this time may be less reliable because of the crude methodology employed (Hayes, 1982; Smith, 2001). The concentrations of DDT in the blood of ordinary people are shown in Table 15.13 in Smith (1991). It is of interest that although each person without special exposure to DDT has relatively constant serum levels of DDT and DDE, DDE values differ more than the DDT values from person to person (Apple et al., 1970). Whether this reflects differences in metabolism or differences in past exposure is unclear. Kreiss and coworkers showed that DDE in serum samples of a community exceptionally exposed to DDT increased with age of the individual (Kreiss et al., 1981). Levels of DDT and its metabolites in the serum from those aged 21 years rose over a 12-month period following application of the pesticide to their homes in KwaZulu, South Africa. In contrast, levels fell in the age group 3–20

Chapter | 93  Toxicology of DDT and Some Analogues

years, showing the complexity of any pharmacokinetic interpretations (Bouwman et al., 1994). Surveys have demonstrated a gradual decline in the concentrations of DDT and related compounds in human fat. Presumably a similar decline has occurred in the levels of these compounds in human serum. Consumption of fish appeared to be a predictor of plasma DDE levels but most reliable were age and serum cholesterol (Laden et al., 1999). When storage of DDT has been found to be greater in black people, the difference could be accounted for by greater exposure (D’Ercole et al., 1976). However, Sandifer who found that the concentrations of DDT in the sera of blacks was two to three times greater than those in whites, also found a significant correlation between total DDT and deficiency of glucose-6-phosphate dehydrogenase, a condition much more common in blacks than whites (Sandifer, 1974). Thus, a genetic factor in the storage of DDT appears possible, but much stronger evidence would be necessary to confirm it. Whether the high storage in blacks is strictly environmental or partly genetic, it is certain that high or higher levels have been recorded among several groups of rural blacks in different parts of the southeastern United States (D’Ercole et al., 1976; Keil et al., 1972a,b) than were reported (Kreiss et al., 1981) among blacks in Alabama, who had mean values of 0.096 and 0.062 ppm for total DDT in the serum of males and females, respectively. Storage of DDT and related compounds in the organs of adults and fetuses in the general population close in time to widespread DDT use was discussed and tabulated by Hayes (1975). Concentrations in the viscera of adults averaged 1.0 ppm, but concentrations in lymph nodes and especially bone marrow (a fatty tissue) approached the level in adipose tissue (6.0 ppm). Concentrations in some viscera of stillborn infants were similar to those in adipose tissue of the same infants and also in adults, suggesting that there had been a mobilization of DDT from fat prior to death. The levels of DDT in human leiomyomatous uterine tissue were significantly higher than those in normal tissue (means of 0.845 and 0.103 ppm, respectively) (Saxena et al., 1987). Whether this is related to any estrogenic actions of DDT is unknown. See also Section 93.2.4.6. Current values of blood DDT levels are widely reported in the literature.

93.2.4.11  Secretion in Milk No information is available on the secretion of DDT in the milk of women who were occupationally exposed to the compound or who were made ill by it, regardless of circumstances. Earlier investigations of the concentrations of DDT in the milk of women in some general populations are shown in Table 15.14 of Smith (1991). Values reported from Guatemala and early values from the USSR were much higher than those from other countries and yet there was no indication of illness among babies fed such milk.

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The significance of DDT in milk and the dosages that different concentrations of it determine have been discussed previously (Coulston, 1985a; Hayes, 1975; Jensen, 1983; Spindler, 1983). Quinby et al. noted that women apparently were in negative DDT balance during lactation, but no direct measurement of DDT intake of women participating in the study was made (Quinby et al., 1965a,b). Subsequently, the ingestion of DDT in food and the secretion of DDT in milk were measured in the same women, and the fact of negative balance was confirmed (Adamovic and Sokic, 1973; Adamovic et al., 1978; Cocisiu, 1976) and may be a significant factor in determining the lower levels of DDT found in women than men in the general population (Adamovic and Sokic, 1973). Significantly lower levels of DDT (mean of 0.008 ppm) and of DDE (mean of 0.035 ppm) had been reported for the milk of city dwellers (Jonsson et al., 1977). However, levels remained quite high (0.05–1.90 ppm) in some rural black people (Woodard et al., 1976). Some evidence suggested that DDT levels are higher in milk from smokers than nonsmokers, although there may be an occupational explanation (Coulston, 1985a). In areas of KwaZulu, human milk levels of DDT and metabolites were significantly higher in women whose houses had been treated with DDT to interrupt malaria transmission (Bouwman et al., 1990a,b). Primiparous mothers had significantly more than multiparous mothers. Transfer from the mother’s milk to the child’s blood was clearly demonstrated (Bouwman et al., 1992). Overall, despite the presence of DDT in human milk and placenta, there seems little risk to neonates in many different populations. Most evidence has shown a continuing decline in DDT levels in humans since it was banned for use in many parts of the world (e.g., Stevens et al., 1993). Clearly, levels would have to be very high before any advice against breast feeding could be given. Recent levels from many parts of the world can be found in the literature, which is continually expanding.

93.2.4.12  Excretion of DDT-Related Compounds Among workers whose DDT intake was estimated to be about 35 mg/day, the concentration of DDA in urine ranged from 0.12 to 7.56 ppm and averaged 1.71 ppm (Ortelee, 1958). Among workers whose exposure was about half as great (Laws et al., 1967), concentrations were from 0.01 to 2.67 ppm with a mean of 0.97 ppm. Continuous sampling of a DDT-formulating plant worker’s urine showed that excretion of DDA increased promptly when exposure began on each of five consecutive workdays but often continued after exposure, sometimes reached a peak about midnight, and then decreased rapidly. On day 6, when there was no occupational exposure to DDT, the excretion of DDA continued until a very low level was reached. The highest concentration of DDA reported in this study was 0.68 ppm (Wolfe

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and Armstrong, 1971). The urine of people in the general population contains not only DDA, but also neutral compounds including p,p-DDT and p,p-DDE (Cueto and Biros, 1967). Men with heavy occupational exposure to DDT excreted much more DDA but showed only a statistically insignificant increase in excretion of DDT and DDE. The urinary excretion of DDT-derived material is of such an order of magnitude that it may account for much of the excretion of absorbed DDT. Examples of the excretion by people of DDA with different kinds and degrees of exposure is presented in Table 15.15 of Smith (1991). It is currently being developed as a method to monitor DDT exposure where the insecticide is currently used in malarial control (B. Krieger, personal communication). DDT and DDE are also excreted in the bile and higher levels than normal were found in the bile of one pest-control operator (Paschal et al., 1974). Further discussion of the storage and excretion of DDT can be found in Levine (1991).

93.2.4.13  Other Laboratory Findings In the absence of occupational DDT poisoning, there has been no opportunity to explore (as has been done with the cyclodiene insecticides) the relationship between clinical and EEG findings. In fact, the only DDT workers studied in this regard were also exposed to benzene hexahydrochloride and benzilan, so the findings might have been related to one or more of the compounds or to their interaction. Electroencephalograms were obtained from 73 of these workers exposed for periods ranging from 7 months to 20 years (Israeli and Mayersdorf, 1973; Mayersdorf and Israeli, 1974). Just over 78% of the records were normal and 21.9% were abnormal. The most severe changes involved persons exposed to the three compounds for 1–2 years; less severe changes were seen with either shorter or longer exposure. The changes were not correlated with age; the range and mean of age for those judged abnormal were almost identical with these values for persons considered normal. Some of the records showed bitemporal sharp waves with shifting lateralization combined with low-voltage theta activity. Other records showed spike complexes, paroxysmal discharges composed of slow and sharp waves most pronounced anteriorly, and low-voltage rhythmic spikes posteriorly. None of the persons examined showed any abnormal clinical neurological finding. The incidence of abnormal electroencephalograms in the general population is 9.0 or 9.2%, according to other investigators cited by Israeli and Mayersdorf (1973); Mayersdorf and Israeli (1974). Some considered that at that time nonspecific EEG abnormalities occurred in 10–20% of the general population (Czegledi-Janko and Avar, 1970), so there is some question of whether the results are meaningful. Clinical laboratory findings associated with DDT poisoning are not specific and it is difficult to diagnose that poisoning has occurred from this agent rather than others.

93.2.4.14  Treatment of Poisoning No useful guidance regarding treatment has been gleaned from the very few cases of DDT poisoning that have occurred. Animal studies indicate that sedatives, ionic calcium, and glucose or another ready source of energy would be useful. On the basis of experience in treating people poisoned by different convulsive poisons, it seems likely that diazepam would be beneficial.

93.3  TDE 93.3.1  Identity, Properties, and Uses TDE is 1,1-dichloro-2,2-bis(4-chlorophenyl)ethane (Figure 93.1). The common name TDE (ISO) is an acronym for tetrachlorodiphenylethane. Except in France, it was a generally recognized name for the compound as a synthetic insecticide. For reasons that are obscure, the word DDD (an acronym for dichlorodiphenyldichloroethane) has been used more commonly for 1,1-dichloro-2,2-bis(4chlorophenyl)ethane when viewed as a metabolite of DDT and this distinction has been retained here. Almost everything we know about this dichloro analogue of DDT relevant to humans is associated with the use of the o, p-isomer as a drug rather than use of the p,p-compound as an insecticide. Nonproprietary names for the o,p-isomer which is used as a drug include chlordithane (USSR) and mitotane (United States). A proprietary name for the insecticide has been Rhothane. Code designations include D-3, ENT-4,225, ME-1,700, and NSC-38,721 (for o,p-isomer only). TDE has the empirical formula C14H10Cl4 and a molecular weight of 320.05. The pure p,p material forms colorless crystals melting at 109–110°C. The technical material consisted mainly of the p,p-isomer but also contained a significant proportion of the o,p-isomer and lesser proportions of related compounds. The insecticidal properties of TDE were first described by Lauger et al. (1944). The formulations have included the technical material; wettable powders, 5%; emulsion concentrates, 25%; and dusts, 5% and 10%.

93.3.2  Toxicity to Laboratory Animals The effects of TDE are similar to those of DDT, but TDE is much less toxic to rats and humans. Gaines found the oral LD50 in both male and female rats to be greater than 4000 mg/kg (Gaines, 1969). Lehman reported 3400 mg/kg as an oral LD50 in rats and 1200 as a dermal value in rabbits (reported in Hayes, 1982). Rabbits were killed quickly by dermal applications at the rate of 400 mg/kg day; they were made severely ill but did not die when treated at the rate of 200 mg/kg/day for 90 days. In rats fed for 2 years, the lowest dietary level producing gross effects was 400 ppm and the lowest level fed (100 ppm, about

Chapter | 93  Toxicology of DDT and Some Analogues

5 mg/kg/day) produced tissue damage. In the rat, pathology is indistinguishable from that caused by DDT (reference details are quoted in Hayes, 1982; Smith, 1991, 2001).

93.3.2.1  Absorption, Distribution, Metabolism, and Excretion The metabolism of p,p-DDD has been described earlier in this chapter. Regardless of dosage form, 75% or more of o,p-DDD is absorbed from the gastrointestinal tract (Korpachev, 1972a,b). Following repeated doses, storage of o,p-DDD reached its highest point in 10–20 days and then decreased somewhat in spite of continued intake. Elimination was rapid after treatment stopped but was detectable longest in the adrenals and adipose tissues. The metabolism of o,p-DDD in the rat has been investigated thoroughly by Reif and Sinsheimer (1975); their major results are included in Figure 93.3, which also records the metabolites found in humans by Reif et al. (1974). More recent studies to explain the covalent binding of o,p-DDD in lung and adrenals are also described Section 93.2.

93.3.2.2  Biochemical Effects The basis for the action of o,p-DDD on the adrenal is not understood fully in connection with any species, but it is clear that marked species differences exist (see Section 93.2). The mechanism that leads to prompt atrophy in the dog may be quite different from the mechanisms that limit the production or increase the breakdown of corticosteroids in species in which most or all of the adrenal cells stay alive. It is clear that a reduction of steroid production accompanies atrophy of the dog. Kupfer considered: (a) reduced steroid production in species other than the dog, including the possibility that such reduction is secondary to inhibition of glucose-6-phosphate dehydrogenase activity in the adrenals, and (b) blockage of steroid action by a steroid metabolite formed under the influence of DDD (Kupfer, 1967). However, the existence of these effects, much less their importance, remains obscure. Hart and Straw showed that administration of o,p-DDD to dogs for only 2–48 h completely blocked the normal increase in steroid production in response to ACTH in vitro but, paradoxically, produced a marked increase in the incorporation of labeled amino acids into protein of the slices (Hart and Straw, 1971c). The same authors presented evidence that the site of action is the intramitochondrial conversion of cholesterol to pregnenolone (Hart and Straw, 1971d), specifically, ACT-activated conversion and not baseline steroid production (Hart and Straw, 1971b). A secondary site involves inhibition of intramitochondrial conversion of 11-deoxycortisol to cortisol. Further evidence supporting the importance of the primary site was also found by Komissarenko et al. (1972).

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o,p-DDD inhibited ACTH-induced steroid production by 797% within 2 h, and the active principle is either o,pDDD per se or a derivative formed in the adrenal gland of the intact dog (Hart and Straw, 1971a). o,p-DDD applied to liver slices in vitro is not effective in reducing ACTHinduced steroidogenesis in the slices. However, the compound did reduce the formation of corticosteroid from progesterone or deoxycorticosterone added to homogenates made from adrenal cortices from dogs, chickens, rats, and human fetuses. These results are consistent with the view that the action of o,p-DDD is to block 11--hydroxylation (Kravchenko, 1973). Furthermore, a concentration of 16 ppm produced this effect in a monolayer culture of human fetal adrenal cells (Komissarenko, 1971). Martz and Straw interpreted the decrease in adrenocortical heme and P450 produced by o,p-DDD in the dog as a suggestion that the compound is metabolized to a more active form (Martz and Straw, 1973), and this is supported by in vitro studies with isolated adrenal mitochondria (Martz and Straw, 1980; Pohland and Counsell, 1985). Whether these actions are related to the covalent binding of 3-methylsulfonyl-p,p-DDE in mouse adrenals by CYP11B has yet to be investigated (see Section 93.2.3.5). There is evidence for a peripheral action of o,p-DDD on steroid transformation in humans also, although the site of action is different. This evidence was obtained by studying the excretion of metabolites of small injected doses of radioactive steroid before and during administration of the drug. It was concluded that 3-hydroxy-5-steroid dehydrogenase was inhibited (Bradlow et al., 1963). Further evidence that o,p-DDD has some inhibitory effect on the synthesis of corticosteroids in humans was provided by in vitro tests on adrenal tissue removed surgically from patients, some of whom had been under treatment with the drug (Touitou et al., 1978). Total doses prior to surgery had varied from 324 to 2280 g and had been given over periods of 1–12 months. Compounds whose synthesis was inhibited in tissue from treated patients were cortisol, corticosterone, 18-hydroxycorticosterone, and aldosterone. Direct addition of o,p-DDD to human adrenal tissue in vitro was without effect on synthesis of corticosteroids. Following massive dosage (60 mg/kg, IV), all of the isomers of DDD inhibit ACTH-induced steroid production in the dog, but the inhibition reached 50% of control in only 27 min after dosing with the m,p-isomer (Hart et al., 1973). There was a marked temporal correlation between the percentage inhibition of ACTH-induced steroid production, the disruption of normal cellular structure of the innermost zones of the adrenal cortex, and the severity of the damage to mitochondria in these zones caused by the three isomers. The effectiveness of m,p-DDT for treating metastatic adrenocortical carcinoma had already been demonstrated (Nichols et al., 1961). However, in humans and dogs, m,p-DDD is less effective than o,p-DDD (deFosse et al., 1968; Reznikov, 1973). Administration of o,p-DDD

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to dogs is followed by a decrease in plasma albumin and an increase in globulins, especially 2-, 1-, and -globulin (Vaniurikhina, 1972). The relation of these changes to the suppression of adrenal function is unknown as is their clinical significance. Guinea pigs receiving o,p-DDD intraperitoneally at a rate of 100, 200, or 300 mg/kg/day for 20 days showed decreases in ascorbic acid levels corresponding to dosage (Petrun and Nikulina, 1970). It was speculated that this might interfere with synthesis of corticosteroids. Like other chlorinated hydrocarbon insecticides, o,p-DDD stimulates hepatic microsomal oxygenation of both drugs and steroids and this may explain much of its action on corticoid metabolism in a wide range of species (Kupfer, 1967). Increased breakdown is evidenced by increased excretion of polar metabolites while nonpolar metabolites remain stable or even decrease – a finding encountered in human patients (Hellman et al., 1973). However, the demonstrated effect on corticoid metabolism fails to explain why o,pand m,p-DDD are unique in their overall effects on the adrenal, including their ability to produce adrenocortical atrophy in the dog. Other powerful inducers of microsomal enzymes lack these effects. Furthermore, in some systems DDD is a relatively weak inducer compared, for example, to DDT and DDE (Gillett et al., 1966). Whereas induction does occur in dogs, its interpretation is complex; for example, the induction caused by repeated doses can be suppressed by cortisol (Martz and Straw, 1972; Mikosha, 1985). Inhibition of NADP reduction by malic enzyme in adrenals may play a role in o,p-DDD action, perhaps by causing a decrease in steroid metabolism (Ojima et al., 1985).

93.3.2.3  Effects on Organs and Tissues DDD is used to control different forms of adrenal overproduction of corticoids in humans. This therapy originally was based on the demonstration that DDD (Nelson and Woodard, 1949) and especially o,p-DDD (Cueto et al., 1958; Komissarenko et al., 1968) cause gross atrophy of the adrenals and degeneration of the cells of its inner cortex in dogs, although it was first thought that DDD produces almost no detectable damage to the adrenals of a variety of species, including humans (Hayes, 1982; Nelson and Woodard, 1949; Smith, 1991, 2001). In the dog, o,pDDT produces gross atrophy of the adrenals when administered at a dosage of only 4 mg/kg/day. The dosage of technical grade DDD required to produce the same effect is 50–200 mg/kg/day (Cueto et al., 1956). However, in spite of its exceptional susceptibility, there is a definite threshold below which the dog does not respond. About 15% of technical DDT was o,p-isomer, much of which is gradually metabolized to o,p-DDD. Yet dogs remained healthy and reproduced normally in a three-generation study involving dosages of technical DDT as high as 10 mg/kg/day.

Hayes’ Handbook of Pesticide Toxicology

DDD has been little used for Cushing’s syndrome in dogs (Lubberink et al., 1971), but it is effective at lower dosages than those used in humans, and side effects are less serious and less frequent (Schechter et al., 1973). It is an interesting fact that p,p-DDE and the -OH analogue of p,p-DDD causes moderate hypertrophy of the dog adrenal and 2,2-bis(p-chorophenyl)ethane causes moderate hyperplasia (Larson et al., 1955). The effect of DDD on thymolymphatic tissues is poorly understood. In one of the earliest studies of the compound, the spleen of all treated animals showed impressive siderosis (Lillie et al., 1947). Much later it was reported that, in rabbits, intramuscular injection of a commercial-grade DDD (mainly p,p-isomer) caused acute atrophy of the thymus and hypertrophy of the adrenal, although the m,p-isomer at a dosage of 100 mg/kg/day caused hypertrophy of the thymus and an increase in its choline acetylase activity (Gawhary, 1972). Decrease in the weight of the thymus and spleen as well as the adrenal glands of rats treated with o,p-DDD was reported (Hamid et al., 1974). Furthermore, Cueto and colleagues (Cueto, 1970; Cueto and Moran, 1968) showed that, at a dosage of 50 mg/kg/day for 14 days, o,p-DDD caused a gradually progressive hypotensive failure in dogs injected with epinephrine or norepinephrine, while leaving unchanged the cardioaccelerator and immediate pressor response of these drugs. The hypotensive failure was associated with weakening of the contractile force of the heart and with a reduction of plasma volume. The latter may have been caused by loss of fluid from the intravascular compartment and was not caused by release of histamine. The hypotensive state could be prevented to a significant degree by pretreatment with prednisolone. In mice, p,p-TDE at a dietary level of 250 ppm moderately increased the incidence of lung tumors in both sexes (Tomatis et al., 1974). DDD (p,p-TDE) is toxic to isolated rabbit Clara cells and human bronchial epithelial cells by what appears to be cytochrome P450 activation to the acyl chloride (Nichols et al., 1995). Leydig cell tumors were reported in the testis of rats receiving o,p-DDD at the rate of 0.6 mg/kg/day for 285–348 days (Lacassagne, 1971). This report is inconsistent with other studies (Hayes, 1982), and this may indicate that a contaminant was involved or a strain variation in the response. In an NCI study, there was a possible effect of TDE in causing an increased incidence of follicular cell carcinoma or follicular cell adenoma of the thyroid in male Osborne-Mendel rats but no effects in female B6C3F1 mice (NCI, 1978a). TDE was found not to be mutagenic in Drosophila (Vogel, 1972). It was found mutagenic in two of three indicator organisms in host-medicated tests but not in direct tests, suggesting that a metabolite was the active agent. However, in the same series of studies, both DDT and DDA were negative (Buselmaier et al., 1973). In addition to atrophy of the zona fasciculata and zona reticularis in the dog, o,p-DDD changes the ultrastructure of most cell

Chapter | 93  Toxicology of DDT and Some Analogues

types of the anterior pituitary of that species. The most striking feature is an increase in corticotrophocytes such as is seen following adrenalectomy, and the increase in cells is presumably associated with increased production of ACTH. The hypothalamus also is involved (Gordienko and Kozyritskii, 1970, 1973). In spite of their severe nature, the changes produced in the dog adrenal are at least partially reversible (Komissarenko et al., 1972). Dosage–response relationships of mitochondrial swelling and of some other details of pathology in the dog adrenal have been explored by Gordienko and Kozyritskii (1973) and by Powers et al. (1974), who also investigated regeneration of the gland. Hypertrophy of the thyroid in dogs receiving 25 mg/kg and its inhibition in those receiving 50 mg/kg had been reported (Gordienko et al., 1972).

93.3.3  Toxicity to Humans 93.3.3.1  Therapeutic Use Following the demonstration that DDD caused atrophy of a part of the adrenal cortex of dogs, the compound has been used in humans in the hope of controlling excessive cortical secretion or of reducing the size of adrenal tumors. The underlying condition may be hyperplasia or adrenocortical carcinoma. Early attempts using mixed isomers and/or dosages less than 100 mg/kg/day often were ineffective, although side effects might be produced (Sheehan et al., 1953). The dosage of o,p-DDD has varied from 7 to 285 mg/kg/day, but a dosage of approximately 40 or more often 100 mg/kg/day for many weeks has been necessary to produce any benefit in humans (Bergenstal et al., 1960; Bledsoe et al., 1964; Gallagher et al., 1962; Gutierrez and Crooke, 1980; Southern et al., 1966a,b; Verdon et al., 1962; Wallace et al., 1961) and other references quoted previously (Hayes, 1982; Smith, 2001). There has been renewed interest recently in adjuvant mitotane therapy following operative treatment for adrenal cortical therapy (Daffara et al., 2008; Igaz et al., 2008; Terzolo and Berruti, 2008). The effects of idiopathic hyperplasia may be controlled; in fact, a state of adrenal insufficiency may be produced (Canlorbe et al., 1971; Helson et al., 1971) or of adrenocortical activity secondary to a tumor that produces ACTH (Carey et al., 1973). Very early attempts to use DDD for treating Cushing’s syndrome often failed because the o,p-isomer was not used and sometimes because the dosage was small, but this was not true for what apparently was the first therapeutic use (Sheehan et al., 1953). Using the o,p-isomer, a favorable response is produced in about one-fourth to one-half of patients with inoperable adrenocortical carcinoma (Canlorbe et al., 1971; Gutierrez and Crooke, 1980; Hoffman and Mattox, 1972; Hutter and Kayhoe, 1966; Lubitz et al., 1973; Montgomery and Struck, 1973). In fact, an occasional cure, involving complete regression of metastases, is produced by

2007

chemotherapy including o,p-DDD (Hoffman and Mattox, 1972; Pellerin et al., 1975; Perevodchikova et al., 1972; Rappaport et al., 1978; Schick, 1973). Other patients have lived for several years (McKiernan et al., 1978). More commonly, symptoms are relieved and life is prolonged only about 7–8 months or a little longer (Canlorbe et al., 1971; Hoffman and Mattox, 1972; Hutter and Kayhoe, 1966; Lubitz et al., 1973) or even less (Hajjar et al., 1975). The success of treatment often is indicated early by a reduction of steroid excretion (Hoffman and Mattox, 1972; Lubitz et al., 1973), but steroid excretion may increase, decrease, or remain unchanged (Fukushima et al., 1971). Removal of the tumor and o,p-DDD treatment may be combined (Levy et al., 1985). The success of treatment is greater in Cushing’s syndrome due to adrenal hyperplasia (Weisenfeld and Goldner, 1962). An early example of what appeared to be complete cure was reported by Bar-Hay et al. (1964). Ten of 17 patients with this condition experienced cure or remission for 12–32 months after the drug had been withdrawn (Luton et al., 1973). The large dosage of o,p-DDD necessary to produce clinical benefit often produces general lassitude, anorexia, nausea, vomiting, diarrhea, and/or dermatitis (Bochner et al., 1969; Danowski et al., 1964; Halmi and Lascari, 1971; Hoffman and Mattox, 1972; Hutter and Kayhoe, 1966; Naruse et al., 1970; Weisenfeld and Goldner, 1962). Gynecomastia, hematuria, leukopenia, and thrombocytopenia have been reported more rarely (Luton et al., 1972; Perevodchikova et al., 1972). The symptoms disappear soon after administration of the drug is stopped or the dosage is reduced. Furthermore, some patients do not develop toxicity. A 10-year-old girl received 7500 mg/ day for a total of 9 kg over 42 months without discernible side effects (Helson et al., 1971). Even large, therapeutic doses of o,p-DDD cause no histological alterations of the adrenals in humans (Wallace et al., 1961). However, electron microscopy revealed degenerative changes in the mitochondria of the zona fasciculata of a patient who had received o,p-DDD at the rate of about 3000 mg/day for 1 month (Temple et al., 1969). Dosages in the therapeutic range (specifically those between 110 and 140 mg/kg/day) produced no detectable injury to the liver, kidney, or bone marrow even though the patients exhibited the reversible symptoms listed earlier (Bergenstal et al., 1960). Kupfer reviewed the extensive literature indicating that the effect in humans and other species, except the dog, is caused by stimulation of corticoid metabolism by massive doses of o,p-DDD and not to any direct reaffect on the adrenal (Kupfer, 1967). Southern et al. (1966a,b) agreed that the effect was predominantly extra-adrenal in humans when the drug was first given but offered evidence that adrenal secretion of cortisol eventually was reduced. Even though therapeutic doses eventually have a direct effect on the adrenal, doses encountered by workers exposed to technical DDT do not (Clifford and Weil, 1972; Morgan and Roan,

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1973). Somewhat encouraging results were reported in the use of p,p-DDD for treating diabetics with hyaline vascular changes and hyperpolysaccharidemia (Tornblom, 1959). Apparently, there has been no attempt to use o,p-DDD for this condition. o,p-DDD has been used, in a much lower dosage, for treating spanomenorrhea associated with hypertrichosis (Klotz et al., 1971). Menstruation was restored in 13 of 15 women with these conditions, and normal pregnancies occurred in five of them during the treatment period. The babies were normal. There was some improvement in hypertrichosis in nine and no improvement in six. At least part of the action of o,p-DDT in controlling excessive androgens involves its action on their metabol­ ism. It was found in a study of three patients with metastatic adrenal carcinoma and one with pernicious anemia that the compound decreased the conversion of androgens to androsterone by about 76% and to etiocholanolone by about 80%. The main effect on androgen metabolism was consistent with induction of microsomal oxidase activity by the drug (Hellman et al., 1973). Uptake of radioactive iodine has been used for diagnosis of Cushing’s syndrome, employing [131I]19-iodocholesterol. DDD labeled with 131I has been used for the same purpose (Skromne-Kadlubik et al., 1974). No comparative study of the duration of storage of the two compounds appears to have been made. However, it is clear that it is possible to introduce enough radiation via 131 I-labeled DDD either to kill rodents or to cause atrophy of their adrenal glands, depending on the schedule of administration (Skromne-Kadlubik et al., 1974). This was viewed as an indication that 131I-labeled DDD might be useful for treating human adrenal carcinoma. In one patient under treatment with o,p-DDD, uptake of radioactivity from [131I]19-iodocholesterol was reduced but not to the normal range (Morita et al., 1972). DDD is commonly found in blood and tissues from the general population. For examples of levels of it in blood, see Table 15.13 in Smith (1991).

93.3.3.2  Analytical Findings Studies associated with what apparently was the first attempt to use p,p-DDD in treating Cushing’s syndrome confirmed that the compound is concentrated in the adrenal gland. Eleven weeks after the last course of DDD, when the concentration in adipose tissue was less than half what it had been earlier, the concentration in an adrenal biopsy was 50 ppm, wet weight. On a lipid basis, the concentrations in fat and adrenal were almost identical (Sheehan et al., 1953). A patient who had received o,p-DDD at the rate of 4000 mg/ day for 58 days had a blood level of 6 ppm and excreted 8.3 mg of free and 39.7 mg of conjugated DDA in a 24-h urine sample (Sinsheimer et al., 1972). There is evidence for two plasma pools of o,p-DDD (Slooten et al., 1982). Normal volunteers excreted increased concentrations of DDA within 24 h of receiving p,p-DDD at a rate of 5 mg/ day and continued to excrete DDA at greater than predose

Hayes’ Handbook of Pesticide Toxicology

levels for over 4 months after dosing was stopped after 81 days (Roan et al., 1971). Treatment of appropriate cases with o,p-DDT usually results in a decrease in urinary steroid excretion (Gutierrez and Crooke, 1980). An unusually detailed study of the individual compounds is that of Hartwig et al. (1968). In long-term administration of o,p-DDD (2 g/day for 1–3 months) to patients with adrenal carcinoma or Cushing’s syndrome, Ojima et al. (1984) found that plasma levels of pregnenolone, progesterone, cortisol, corticosterone, and some other C21 steroids were progressively decreased, as well as urinary excretion of 17-ketosteroids and 17-hydroxycorticosteroids. Touitou et al., however, were unable to demonstrate any correlation between concentrations of o,p-DDD in adrenals removed from patients preoperatively treated with the drug for Cushing’s syndrome and inhibition of some steroid biosynthesis enzymes measured in vitro (Touitou et al., 1985). There is a suggestion that o,p-DDD suppresses ACTH-secreting cells in the pituitary as well as depressing steroid hormone secretion (Takamatsu et al., 1981). In some patients, either p,pDDD (Tornblom, 1959) or o,p-DDD (Molnar et al., 1961) caused an increase in plasma cholesterol, but the opposite may also occur (Danowski et al., 1964). Oddly enough, such patients are refractory to the therapeutic effects of the drug (Molnar et al., 1961). o,p-DDD may have a hypouricemic effect apparently by increasing the renal clearance of uric acid (Reach et al., 1978; Zumoff, 1979).

93.4  Ethylan 93.4.1  Identity, Properties, and Uses Ethylan is 1,1-dichloro-2,2-bis(4-ethylphenyl)ethane, that is, the p,p-diethyl analogue of DDD (Figure 93.1). Ethylan is the only common name in use, but apparently was approved officially only in the USSR. The trade name Perthane was often used. Code designations for ethylan include B-63,138 and Q-137. The CAS registry number is 72-56-0. Ethylan has the empirical formula C18H20Cl2 and a molecular weight of 307.27. The pure compound is a crystalline solid with a melting point of 56–57°C. The technical product is a waxy solid with a melting point not below 40°C and with some decomposition above 52°C. The insecticide is practically insoluble in water but soluble in acetone, kerosene, and other organic solvents. Ethylan was introduced in 1950 and has been used to control pear psylla, leaf hoppers, various larvae on vegetables, and moths and carpet beetle on textiles.

93.4.2  Toxicity to Laboratory Animals The oral LD50 values for ethylan were 8170 and 9340 mg/kg in rats and mice, respectively. However, the corresponding intravenous values were only 73 and 173 mg/kg in the

Chapter | 93  Toxicology of DDT and Some Analogues

same species. No dermal LD50 value could be measured; all rabbits that received a 30% solution at the rate of 3 ml/ kg/day for 13 weeks survived. Finnegan et al. (1955) and Gaines (1960) agreed that the oral toxicity was very low (LD50 > 4000 mg/kg). Minimal and infrequent changes were seen in the livers of rats fed dietary levels of 2500 and 5000 ppm for 2 years. There was no effect on survival, and differences in growth rate did not correspond to dosage. Thus a dietary level of 1000 ppm might be considered a no-effect level. In contrast, the same investigators found that a dietary level of 5000 ppm was lethal to dogs within 22 weeks. Levels of 100 or 1000 ppm did not interfere with survival or growth when fed for 1 year, although the 1000 ppm level led to some atrophy of the adrenals (Finnegan et al., 1955). Cortisone given at the same time as ethylan tended to block the effect of the latter on the adrenal (Bleiberg and Larson, 1957). Reznikov considered the action of p,p-ethylan similar to that of p,p-DDD (Reznikov, 1973).

93.4.2.1  Absorption, Distribution, Metabolism, and Excretion Rats fed ethylan at a concentration of 50 ppm (about 2.5 mg/ kg/day) for 6 weeks stored the compound in their fat at a concentration of 19 ppm (Finnegan et al., 1955). Four generations of rats were fed a standard synthetic diet containing 20% fat to which several pesticides were added. The diets of the seven groups studied differed only in the kinds of fat (cottonseed oil, lard, etc.) they contained. The average concentration of added ethylan found by analysis in different samples of dietary fat varied from 2.01 to 2.71 ppm. No ethylan was detectable in the body fat or other tissues of the rats (Adams et al., 1974). After a single dose of [14C] ethylan, rats excreted 90% of the radioactivity in their feces and 5% in their urine in 2 weeks (Bleiberg and Larson, 1957).

93.4.2.2  Biochemical Effects Ethylan reduced excretion of 17-hydroxycorticosteroids and caused adrenal atrophy in dogs (Cobey et al., 1956). Dogs that had received ethylan for 10 or 14 days at a rate of 200 mg/kg/day, slept 12–14 h following anesthesia with sodium pentobarbital compared to only 6–7 h following the same dosage of barbiturate before receiving ethylan. Similar studies with different DDD formulations revealed that increased sleeping time did not depend on the presence of adrenal atrophy, but they did not exclude the possibility that it depended on altered function of the adrenal. The increased sleeping time did not depend on reduced clearance of the barbiturate from the blood, which was not influenced by ethylan or DDD. Thus the cause remained obscure. Whatever the cause, the increase in sleeping time was peculiar to dogs and did not occur in rats treated with DDD (Nichols

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et al., 1958). In dogs, ethylan (50 mg/kg/day for 10 days) significantly increased the glutathione reductase of the adrenal cortex but not of the liver (Komissarenko et al., 1978).

93.4.2.3  Effects on Organs and Tissues Ethylan produces adrenal cortical atrophy in the dog (Finnegan et al., 1955; Larson et al., 1955). No such effect was noted in the rat. Presumably the effect is virtually specific for the dog, as is true of o,p-DDD, which has been more extensively studied. Dogs killed or rendered moribund by dietary level of 5000 ppm (about 105 mg/kg/day) showed marked atrophy of the adrenal cortex; the medulla was unaffected. The capsule was wrinkled, the zona glomerulosa contained cells with granular and diminished cytoplasm, and there was a focal loss of these cells. The zona fasciculata was greatly narrowed, and there was extreme vacuolization among cells in the medial two-thirds. However, fat-staining material was deficient throughout the fasciculata, especially the inner part. The zona reticularis had practically disappeared, leaving only a few cells containing lipochrome pigment. There were a few focal concentrations of lymphocytes. Atrophy was present but less severe in two of three dogs that received 1000 ppm (about 21 mg/kg/day). On the contrary, severe atrophy was produced in less than 3 weeks by an oral dosage of 200 mg/kg/day. When ethylan was administered to mice at the highest tolerated rate for about 18 months, the results for tumorigeni­city were equivocal (Innes et al., 1969). Some evidence for hepatic tumor formation in female mice but not males or rats of either sex has been reported (NCI, 1979).

93.4.3  Toxicity to Humans Ethylan was administered to nine men with metastatic carcinoma of the prostate and to five women with metastatic carcinoma of the breast because there had been reports of a favorable effect of surgical adrenalectomy on the clinical course of some patients with these diseases and because the compound had been shown to cause adrenocortical atrophy in the dog (Taliaferro and Leone, 1957). All the patients also received ACTH either intermittently or continuously. With one exception, the dosage of ethylan varied from 50 to 150 mg/kg/day, the latter for a total of 189,000 mg within 21 days. The most intensive treatment was 200–300 mg/kg/day for a total of 96,000 mg in 6 days. The smallest dosage produced diarrhea, vomiting, and especially nausea in some patients and required cessation of treatment. In contrast, other patients, especially those who were less sick to begin with, tolerated the higher dosages, including 200–300 mg/ kg/day, with no symptoms whatever. Marked thrombocytopenia and leukopenia were noted in one patient just after a 14-day course of treatment. These changes, which

2010

were attributed to ethylan, resolved promptly when treatment was stopped. There was no other evidence of hemato­ poietic toxicity and no evidence of hepatic, renal, neural, or other toxicity. It was not considered that the relatively brief treatments influenced the clinical course in any of the cases. However, among patients receiving 150–300 mg/ kg/day in divided doses, ethylan caused a marked depression of plasma 17-hydroxycorticosteroid levels, but never below the normal range. Lower dosages had no consistent effects. No distinct benefit but nausea, vomiting, and skin rash were seen in four other patients with carcinoma who received ethylan at dosages of 1800–8000 mg/day for periods of 4–54 days (Weisenfeld and Goldner, 1962). In surveys of workers and the general population who have developed pancreatic cancer, association with ethylan exposure has been reported (Garabrant et al., 1992, 1993). However, those studies involved self-reporting exposure. See also Section 93.2.4.7.

93.5  Methoxychlor 93.5.1  Identity, Properties, and Uses Methoxychlor is 1,1,1-trichloro-2,2-bis(4-methoxyphenyl) ethane or 2,2-bis(p-methoxyphenyl)-1,1,1-trichloroethane, that is, the p,p-dimethoxy analogue of p,p-DDT. The structure is shown in Figure 93.1. The common name, methoxychlor (BSI, ICPC, ISO), is in general use. Other nonproprietary names have included dianisyl trichloroethane, dimethoxy-DT, DMDT (an acronym for dimethoxydiphenyltrichloroethane), and methoxy DDT. One trade name is Maralate. Code designations include OMS-466. The CAS registry number is 72-43-5. Methoxychlor has the empirical formula C16H15Cl3O2 and a molecular weight of 345.65. The pure material forms colorless crystals melting at 89°C. Technical methoxychlor is a gray flaky powder containing about 88% of the p,p-isomer, the remainder being mainly o,p-isomer, although up to 50 other contaminants have been detected (Lamoureux and Feil, 1980; West et al., 1982). The density is 1.41 at 25°C. Methoxychlor is stable to heat and ultraviolet light and resistant to oxidation; it is dehydrochlorinated by alkalies and by heavy metal catalysts. The solubilities of methoxychlor are approximately the same as those of DDT. It is readily soluble in most aromatic solvents, moderately soluble in alcohols and petroleum oils, and essentially insoluble in water. Methoxychlor was first described by Lauger et al. (1944) and it was introduced about 1945. Formulations include wet-table powder (25 and 50%), emulsifiable concentrate (24%), dusts (4–10%), and aerosols. Methoxychlor is effective against a wide range of insects affecting fruits, vegetables, forage crops, and livestock. The low toxicity of methoxychlor and its short biological half-life were largely responsible for its greatly expanded use following the ban on DDT in many countries.

Hayes’ Handbook of Pesticide Toxicology

93.5.2  Toxicity to Laboratory Animals 93.5.2.1  Basic Findings In spite of its low toxicity, methoxychlor in sufficient dosage is capable of causing convulsions in the dog (Tegeris et al., 1966). However, in the rat, the compound causes depression of the central nervous system (Lehman, 1951–52 quoted in Hayes, 1982; Smith, 1991). Tremors have been noted, but they are not a prominent symptom. In rabbits killed by a few doses, the only signs noted were diarrhea and anorexia (Smith et al., 1946). The acute toxicity of methoxychlor is very low; oral LD50 values of 5000 mg/kg (Hodge et al., 1950) and 6000 mg/kg (Hayes, 1982) have been reported for the rat and values of 1850 mg/kg (Domenjoz, 1946) for the mouse and 2000 mg/kg for the hamster (Cabral et al., 1979). Rats fed a dietary level of 30,000 ppm suffered a severe reduction in growth, and most of them died in less than 45 days. Those fed a dietary level of 10,000 ppm survived but gained almost no weight; paired feeding studies showed that failure of growth was due entirely to food refusal. A dietary level of 1600 ppm for 2 years caused measurable reduction of growth but produced no reduction in life span and no histological change in the tissues. Dosages of 20 mg/kg/day for 1 year or a dietary level of 200 ppm (about 10 mg/kg/day) for 2 years were both no-effect levels (Hodge et al., 1950, 1952). Rabbits are relatively susceptible to methoxychlor; a dosage of 200 mg/kg/day was fatal within 15 days (Smith et al., 1946). Dogs fed the compound in their diet in such a way that they received 1000 mg/kg/day for 6 months lost weight, and many of those that received 2000, 2500, or 4000 mg/kg/day began having convulsions within 6 weeks and died within 3 additional weeks. Strangely enough, dogs that received 2500 mg/kg/day administered by gastric tube as a suspension in 1% gum tragacanth for 5 months showed no indication of injury, although the amount of absorption was not clear (Tegeris et al., 1966, 1968). In a study lasting 1 year, the highest dosage fed to dogs (300 mg/kg/day) caused weight loss for a month, but later the dogs on the two lower dosages regained their original weight (Tegeris et al., 1966). A dietary level of 2500 ppm for 8–16 weeks produced no ill effects in chickens (Lillie et al., 1973).

93.5.2.2  Absorption, Distribution, Metabolism, and Excretion Mixtures containing methoxychlor appeared to pass directly from the luminal border of the rat jejunum to the intercellular spaces immediately, in addition to the usual transport through the endoplasmic reticulum toward the Golgi apparatus. Furthermore, absorption of methoxychlor was accompanied by distention of vesicles and intracellular spaces (Imai and Coulston, 1968). Storage of methoxychlor was minimal in the fat of rats that had received it for 18 weeks. No storage could be measured at a dietary level of 25 ppm,

Chapter | 93  Toxicology of DDT and Some Analogues

and storage decreased after the 9th week at levels of 100 and 500 ppm in spite of continued intake. Two weeks after dosing was discontinued, no methoxychlor was detected (Kunze et al., 1950). Storage of methoxychlor in sheep reaches a steady state in 6–8 weeks or, at certain dosage levels, actually declines after that interval in spite of continued intake, perhaps due to induced metabolism. Storage loss is prompt after dosing is stopped (Reynolds et al., 1975, 1976). Following oral administration of radioactive methoxychlor to mice, 98.3% was recovered from the excreta within 24 h. The compound was metabolized to 2-(p-hydroxyphenyl)2-(methoxyphenyl)-1,1,1-trichloroethane and 2,2-bis(phydroxyphenyl)-1,1,1-trichloroethane, which were eliminated largely in conjugated form (Kapoor et al., 1970). Detailed studies have also been conducted with goats (Davison et al., 1982, 1983). The normally low rate of storage of methoxychlor was not influenced by simultaneous feeding of DDT or dieldrin (Street and Blau, 1966). Rat liver microsomes form both the mono – and dihydroxy products as well as more polar compounds (Bulger et al., 1985; Ousterhout et al., 1981). With (R)- and (S)-[monomethyl-2H3]methoxychlors, intramolecular deuterium effects and enantiotopic differentiation have been observed and the demethylations appeared due to CYP2B2 and CYP2C6 isoforms (Ichinose and Kurihara, 1987; Kishimoto and Kurihara, 1996; Kishimoto et al., 1995). A great deal of the metabolism work has been done by Kupfer and colleagues using microsomes, purified cytochrome P450 isoforms, or insect cells containing human cytochrome P450 species (Dehal and Kupfer, 1994; Stresser and Kupfer, 1997, 1998a,b). In the rat, CYP2B isoforms are not only involved in demethylation, but also actively hydroxylate in the ortho positions of methoxychlor and the mono- and dimethoxy analogues. In humans, CYP2C19 and CYP1A2 seem to be responsible for demethylation. Ortho hydroxylation of the monodemethylated products is predominantly catalyzed by CYP3A4. On the whole, bidemethylation and ortho hydroxylation seem to occur less readily with human samples than with rat liver. Another route of metabolism is an initial dechlorination before replacement of the methoxy groups (Bulger et al., 1983; Davison et al., 1983; Kupfer et al., 1986). Phenolic products from both routes appear to be converted to activated intermediates, which may bind covalently to macromolecules (Bulger et al., 1983; Kupfer et al., 1986). This might be related to the formation of the catechols formed by ortho hydroxylation (Bulger and Kupfer, 1989; Kupfer et al., 1990). A simplified scheme of methoxychlor metabolism is shown in Figure 93.4. More recently, the metabolism of methoxychlor by cDNA expressed human cytochrome P450 protein has been explored in detail and 1,1,1,-trichloro-2-(4-methoxyphenyl)2-(3,4-dihydroxyphenyl)ethane identified as a key intermediate (Hu and Kupfer, 2002). Not surprisingly, comparisons of in vitro hepatic metabolism of methoxychlor show species variation and sexual dimorphism of phase I and phase II metabolites (Ohyama et al., 2004, 2005).

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93.5.2.3  Biochemical Effects Methoxychlor induced liver microsomal enzymes in sheep, but even at a dietary level of 2500 ppm the degree of induction was less than that caused by DDT at a dietary level of 250 ppm. Methoxychlor caused no change in food consumption, body weight, liver weight, uterine weight, or estrous cycle (Cecil et al., 1975). In rats, methoxychlor induced hepatic CYP2B1 and CYP3A enzyme levels, but in regimens of multiple treatment and less efficiently than DDT (Li et al., 1995). Further studies have demonstrated a sexual dimorphism in CYP enzymes induced by methoxychlor (Oropeza-Hernandez et al., 2003). Some increases in hepatic glucose-6-phospatase activity have been observed in rats given methoxychlor without showing any histological damage (Morgan and Hickenbottom, 1979) and significant decreases in lactate levels occurred even at doses less than 1% of the oral LD50. Administration of methoxychlor to neonatal rats (Lamartiniere et al., 1982) resulted in elevated levels of sexmonoamine oxidase activities in adult rats, implying changes in the brain hormone environment during development that did not become apparent until adulthood (Lamartiniere et al., 1982). Whereas massive doses of methoxychlor have an estrogenic effect in swine (Tegeris et al., 1966) and perhaps other species, no such effect is detectable in chickens at a dietary level of 10 ppm (Foster, 1973) or in sheep at a dietary level of 2500 ppm (Cecil et al., 1975). Even 10 ppm is a far greater residue than humans or livestock are likely to encounter. Microsomal metabolism of methoxychlor in rat liver has been reported to result in binding to iodothyronine 5-monodeiodinase at cysteine or lysine residues with resulting depression of iodinase activity in vivo. The significance of these findings on thyroid hormone metabolism and action is unclear (Zhou et al., 1995).

93.5.2.4  Effects on Organs and Tissues Of several polycyclic aromatic compounds said to be impurities in commercial methoxychlor, only one was mutagenic and in only one strain (Grant et al., 1976). Mixed mutagenicity results have been more recently reported for methoxychlor (Oberly et al., 1993). Some tumors were found in rats fed for 2 years at dietary concentrations as high as 1600 ppm, but the kind and incidence did not differ from those in controls. No tumors were found in dogs that had received dietary levels up to 10,000 ppm (Hodge et al., 1952). Negative results also were found in mice that received a single subcutaneous injection (10 mg/mouse) or in others given weekly skin applications (0.1 or 10 mg). When two strains of mice were fed technical methoxychlor at a dietary level of 750 ppm for up to 2 years, the incidence and malignancy of carcinoma of the testis were increased in one strain but not the other. It was suggested that the carcinogenicity was related to the estrogenic activity of methoxychlor (Reuber, 1979b). The occurrence of neoplasms of all sorts showed a very rough dosage response in male rats fed technical methoxychlor at

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CCl3 Dechlorination + oxidative metabolism CH3O

OCH3

CCl3 CCl3 HO

HO

OCH3

HO

OCH3

CCl3

HO

OH

CCl3 HO HO

Covalent binding of semiquinone radical OH

Figure 93.4  Main metabolic routes of methoxychlor. The exact mechanisms of covalent binding are not known.

dietary levels of 100 ppm or more and in females fed 10 ppm or more; tumors identified as carcinomas were found in both sexes but only in rats fed 2000 ppm for 2 years (Reuber, 1979a). The conclusion that methoxychlor is carcinogenic for the liver in C3 H and BALB/c mice and Osborne-Mendel rats was proposed as being related to the covalent binding of activated metabolites (Kupfer and Bulger, 1987). However, the original reports, including that from the National Cancer Institute (NCI, 1978b) show only poor evidence for carcinogenicity of methoxychlor. At a dietary level of 1000 ppm, the tumorigenic property of methoxychlor was less than additive when it was fed in combination with Aramite, DDT, and thiourea or Aramite, DDT, and aldrin (Deichmann et al., 1967). Large doses of methoxychlor (1000, 2000, or 4000 mg/ kg/day) produce dosage-dependent chronic nephritis and hypertrophy of the kidneys, mammary glands, and uteri of swine (Tegeris et al., 1966). In rats, Hodge could not find characteristic liver cell changes at any dosage level (Hodge et al., 1950). Striking testicular atrophy at dietary levels of 10,000 ppm or greater was observed and this was not present in parfed controls. Similar findings were reported by Tullner and Edgcomb (1962). No atrophy was present in a dog that had received methoxychlor at 100 mg/kg/day for 1 month. Under certain conditions, continued massive dosage of methoxychlor produced cystic tubular nephropathy in rats (Tullner and Edgcomb, 1962). Wistar rats given methoxychlor (100 or 200 mg/kg/day) for 760 days (males) or 14 days (females) showed inhibition of spermatogenesis and folliculogenesis (Bal, 1984). There

were degenerative changes in Sertoli cells and in spermatogonia and spermatocytes, with some transformed to polynucleate cells. Cytoplasmic vacuolations were observed in the epithelium of the ductus epididymis. In mice at least, the transporter ABCC1 (MRP1) seems to afford some protection of the seminiferous epithelium by mediating the efflux of the pesticide (Tribull et al., 2003). Atresia of the ovarian follicles was evident in rats, with pyknosis and karyorrhexis of the granulosa cells (Bal, 1984) possibly by the Bcl-2 and Bax-mediated pathways (Miller et al., 2005). Recently, oxidative damage and dysfunction of mouse ovary and brain mitochondria have been observed to be caused by methoxychlor (Gupta et al., 2006; Schuh et al., 2005) with inhibition of steroidogenesis in vitro by the metabolite 2,2-bis(p-hydroxyphenyl)-1,1,1,trichloroethane (Miller et al., 2006; Zachow and Uzumcu, 2006). Proliferation of ovarian surface epithelium was also stimulated (Borgeest et al., 2002; Symonds et al., 2005). In rats, Lehman found that the lowest dietary level producing tissue damage was 550 ppm (about 25 mg/kg/day); the effects, which were confined to the liver, consisted mainly of a slight increase in incidence of hepatic cell adenomas (quoted in Hayes, 1982; Smith, 1991). In rats and monkeys, the early induction of liver enzymes and the concomitant increase in hepatic endoplasmic reticulum may be temporary, disappearing in animals treated with methoxychlor for a prolonged time (Serrone et al., 1965). In dogs, a dosage of 2500 mg/kg/day caused grossly visible congestion of the intestinal mucosa. It has also caused progressive degeneration of the mitochondria of mucosal cells of the small intestine marked in the

Chapter | 93  Toxicology of DDT and Some Analogues

early stages by matrical swelling and disruption of the cristae and later by disappearance of cristae and appearance of small myelin bodies. The mitochondria of these cells showed some recovery in a dog that had been returned to uncontaminated food for only 3 weeks after 12 weeks of the high dosage of methoychlor (Tegeris et al., 1968). For other effects on organs, see the two following sections on reproduction.

93.5.2.5  Effects on Reproduction Large doses of methoxychlor have estrogenic effects (Tullner, 1961). Methoxychlor levels of 2500 and 5000 ppm reduced mating, and only one litter was produced. However, when the same rats were returned to an uncontaminated diet, they reproduced normally. A dietary level of 1000 ppm started before mating and continued throughout lactation had no effect on reproduction in that generation, but the female pups had early vaginal opening and reduced reproduction when mature, and the reproductive behavior of mature male pups was also defective (Harris et al., 1974). The potency appeared to be 1/10,000 of diethylstilboestrol. When male and female weanling rats were dosed with methoxychlor at 100 mg or 200 mg/day through puberty and gestation until day 15 of lactation in females (Gray et al., 1989), various parameters of reproductive potential were altered in both sexes, but fertility was only reduced in females when mated with untreated animals. Puberty was delayed in male and female rats after feeding 1200 ppm methoxychlor from gestational day 15 to postnatal day 10 (Masutomi et al., 2003). No implantations were observed in another study when female rats received the same doses of methoxychlor for 14 days before mating and during pregnancy (Bal, 1984). Preimplantation effects of methoxychlor seemed more important than postimplantation in ensuring success (Cummings and Gray, 1989). In vitro tests showed that pure methoxychlor itself is not estrogenic, although the commercial product had some activity and bis(4hydroxyphenyl)trichloroethane was quite active (Bulger and Kupfer, 1977; Bulger et al., 1978a,b). The estrogenic activity of impure methoxychlor in inducing uterus growth, uterine orinithine decarboxylase and epidermal growth factor receptor, creatine kinase and peroxidase (Bulger et al., 1978a; Cummings and Metcalf, 1994, 1995; Metcalf et al., 1995) appears to be caused by the demethylated analogues (Bulger et al., 1985; Ousterhout et al., 1981), which are also metabol­ ites (see Section 93.5.2.2). By both in vitro and in vivo criteria, 1,1-dichloro bis(4-hydroxyphenyl)ethene is the most potent agent (Bulger et al., 1985; Cummings, 1997; Kupfer and Bulger, 1987). The estrogenic effects of methoxychlor are not restricted to those on uterine physiology and function. Both running wheel activity (estrogen controlled) and sex behavior in rats and hamsters were induced by 400 mg/kg/day (Gray et al., 1988). The actions of methoxychlor were not, however, completely identical to those of estradiol. Exposure of pregnant mice to methoxychlor has been reported to cause changes in behavior of male offspring (Vom Saal et al., 1995).

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Methoxychlor affects the decidual cell response of the rat uterus (Cummings and Gray, 1987, 1989), a technique mimicking the growth and development of the endometrium during pregnancy by a mechanism that apparently occurs by interaction directly with the uterus, probably involving disruption of estrogen receptor binding to the estrogen response element of HOXA10, and other genes, (Fei et al., 2005; Waters et al., 2001). Other mechanisms such as embryo transport rate might also be involved (Cummings and Perreault, 1990). Some evidence suggests that reproductive effects of methoxychlor metabolites in male rats (Bal, 1984; Tullner and Edgcomb, 1962) may be mediated, in part, by elevation of prolactin concentration and release, which in turn influences hypothalamic levels of gonadotropin-releasing hormone (Goldman et al., 1986). In studies of the effect of methoxychlor on reproductive tract development following neonatal exposure of mice, precocious vaginal opening, cornification and increased tract size, and ovarian atrophy were observed in females and reduced serum testosterone, testicular DNA content, seminal vesicles, and prostate in males (Cooke and Eroschenko, 1990; Eroschenko and Cooke, 1990; Eroschenko et al., 1995). Changes in females were not, however, completely identical to those observed with 17-estradiol (Eroschenko, 1991). On the other hand, uterine luminal proteins were identical following technical methoxychlor and 17-estradiol administration to ovariectomized mice (Rourke et al., 1991), as were morphometric parameters (Swartz et al., 1994), although some toxicity was observed. The effects of 2,2-bis(p-hydroxyphenyl)-1,1,1-trichloroethane on testosterone formation by Leydig cells from rats was shown not to be due to estrogen receptor binding or to antiandrogenic effects but directly on cholesterol side-chain cleavage (Murono and Derk, 2004). When rats received methoxychlor intragastrically on days 6–15 of pregnancy, dosages of 50–400 mg/kg/day reduced maternal weight gain (Khera et al., 1978). At 200 mg/kg/day, the compound decreased the number and weight of fetuses and caused delayed ossification leading to wavy ribs and other bent bones. No real teratogenesis was observed and no effects were observed in vitro or human and rat testicular cells when examined for single-strand DNA breaks (Bjorge et al., 1996; Khera et al., 1978). Methoxychlor prevented ovariectomy-induced bone loss in the rat (Dodge et al., 1996). Chapin et al. (1997) dosed rats before and following birth and looked for immune and reproductive changes at doses of 0, 5, 50, or 150 mg/kg/day in a large study. Primary adult effects were reproductive and 5 mg/kg/day was not a NOAEL (no observable adverse effect level). A predictable result of the rapid metabolism and excretion of methoxychlor is the fact that very little of it is excreted in the milk and is thus of low risk to humans. When the compound was fed to cows at a dietary concentration of 7000 ppm, the concentration in the milk reached slightly over

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2 ppm in 91 days and remained essentially constant until feeding was stopped on day 112 (Gannon, 1959). After dosing was stopped, the concentration in milk fell to less than 0.1 ppm in a week. In mice, methoxychlor fed to lactating dams did affect the reproductive tract of suckling females (Appel and Eroschenko, 1992). In chickens, dietary levels as high as 2500 ppm (about 145 mg/kg/day) had no effect on the health of hens or their production of hatchable eggs or on the ability of cockerels to fertilize the eggs. The hens were tested for 16 weeks and the cockerels for 8 weeks (Lillie et al., 1973). In summary, methoxychlor possesses many of the estrogenic properties of 17-estradiol probably after in vivo demethylation. However, fast metabolism and low potency (> 1/10,000 of 17-estradiol) pose questions as to how high the risk is to humans. Hall et al. suggested that in mice methoxychlor acts as an estrogen agonist in the uterus but as antagonist in the ovary, suggesting that its reproductive toxicity properties are not identical to estrogen (Hall et al., 1997).

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slightly dizzy. On day 5, rapidly progressive renal failure required hemodialysis. Soon afterward, peripheral sensory and motor neuropathy appeared, including hypoesthesias, paresthesias, persistent leg and foot pain, bilateral footdrop, and difficulty in moving the extremities. A generalized rash appeared. In spite of marked recovery, the patient still had profound, bilateral, sensorineural hearing loss, tinnitus, and moderate neuropathy of the legs and arms when he was reevaluated over 6 years later (Harell et al., 1978). The delay in onset and the character of the illness were not consistent with poisoning by methoxychlor, malathion, or a combination of them, regardless of dosage. No thought seems to have been given to other possible causes, whether toxic or otherwise. A 49-year-old man who was exposed to a dust of methoxychlor and captan developed aplastic anemia a few weeks later and died within 6 months. He had also had light exposure to methoxychlor during the previous 2 years without symptoms (Ziem, 1982).

93.5.3.3  Laboratory Findings

93.5.3  Toxicity to Humans 93.5.3.1  Experimental Exposure Groups of volunteers were given methoxychlor at rates of 0, 0.5, 1, and 2 mg/kg/day for 8 weeks. Even the highest dosage was without detectable effect on health, clinical chemistry, or the morphology of blood, bone marrow, liver, small intestine, or testis (Stein et al., 1965). The highest dosage administered by Stein was similar to 1.4 mg/kg/day, which is considered safe for occupational intake, as reflected in the threshold limit value of 10 mg/m3. The low sensitizing property of methoxychlor has been noted (Szarmach and Poniecka, 1973).

93.5.3.2  Use Experience There apparently has been no confirmed case of poisoning, occupational or otherwise, involving methoxychlor alone. However, atypical cases associated with methoxychlor have been reported. A 21-year-old man first noticed symptoms 8–9 h after spraying several fruit trees with a formulation diluted from a mixture containing methoxychlor (15%) and malathion (7.5%) (Harell et al., 1978). The entire task took only 15–20 min, and afterward a shower was taken. The first symptoms were blurring of vision and gradual onset of nausea. Next morning, the man began vomiting and developed severe diarrhea. He sought medical help 24 h after exposure and was admitted to hospital about 36 h after exposure. He was then dehydrated and suffering severe abdominal cramps and continuing diarrhea. On day 3 after exposure, jaundice was noted and continuous bilateral tinnitus began. On day 4, the patient was completely deaf and

Most investigators have not found methoxychlor in human tissue. Apparently, the first exceptions were Griffith and Blanke who reported finding the compound infrequently and in unstated concentrations in blood taken at autopsy under the medical examiner system of Virginia (Griffith and Blanke, 1975). Under the circumstances of collection, the possibility of occupational exposure of the deceased could not be excluded. The reported persistence of methoxychlor for at least 7 days on the hands of a worker is interesting in view of the rapid metabolism of the compound once it is absorbed (Kazen et al., 1974). However, it must be said that with modern analytical sensitivities, methoxychlor or metabolite residues might be detected more frequently.

93.5.3.4  Treatment of Poisoning In the unlikely event that treatment is required, it must be symptomatic.

93.6  Chlorobenzilate 93.6.1  Identity, Properties, and Uses The IUPAC name for chlorobenzilate (BS1, ISO, JMAF) is ethyl 4,4-dichlorobenzilate. Other names are 4,4-chlorobenzilic acid ethyl ester, ethyl 2-hydroxy-2,2-bis(3-chlorophenyl) acetate, and ethyl 4,4-dichlorodiphenyl glycolate. For the structure see Figure 93.1. Among many proprietary names have been G23992, Acaraben, Benz O-chlor, Benzilan, and Kop-Mite. The CAS registry number is 51015-6. Chlorobenzilate has the empirical formula C16H14Cl2O3 and a molecular weight of 325.20. It is a colorless

Chapter | 93  Toxicology of DDT and Some Analogues

solid melting at 37–37°C. It is very soluble in acetone and hexane but virtually insoluble in water. Impurities in the technical product, which is about 95% pure, can be dichloro­ benzophenon, the ethyl ether of chlorobenzilate, and 4,4-dichlorobenzil. Chlorobenzilate was introduced as a technical product in 1952. It has been used mainly as a miticide on citrus crops or to control mites in beehives.

93.6.2  Toxicity to Laboratory Animals The acute oral LD50 to mice, rats (Horn et al., 1955), and hamsters is about 700 mg/kg. Symptoms in rats and mice include depressed motor activity and rapid wheezing respiration. Dogs tolerated daily oral doses of 64 mg/kg for 35 weeks and rats 500 ppm in the diet for 2 years (Horn et al., 1955). After daily chlorobenzilate doses of 12.8 mg/kg to dogs, 5 days/week for 35 weeks, approximately 40% of the total dose was excreted unchanged or as urinary metabolites. No significant storage in fat of dogs or rats was reported (Horn et al., 1955). Knowles and Ahmad described the conversion of chlorobenzilate by rat liver homogenates to p,p-dichlorobenzilic acid, p,p-dichlorbenzophenone, p,p-dichlorobenzyhydrol, and p-chlorobenzoic acid (Knowles and Ahmad, 1971). In carcinogenicity studies chlorobenzilate induced hepatocellular carcinomas in mice, but the evidence in rats is uncertain (NCI, 1978c). Some testicular atrophy was observed in rats.

93.6.3  Toxicity to Humans A case of a pesticide sprayer poisoned by chlorobenzilate has been described (Ravindran, 1978). Symptoms included ataxia, delirium, fever, and muscle pains. Chlorobenzilate was detected in the urine of some workers employed in Florida citrus groves. Exposed workers had levels ranging from 0.07 to 6.2 mg/l. It should be noted that the methodology employed involved oxidation to p,p-dichlorobenzophenone and would not distinguish between the parent chemical and some of its metabolites (Levy et al., 1981).

93.7  Dicofol 93.7.1  Identity, Properties, and Uses The IUPAC name is 2,2,2-trichloro-1,1-bis (4-chlorophenyl)2,2,2-trichloroethanol, or 1,1-bis(p-chlorophenyl)-2,2,2-trichloroethanol, or 4,4-dichloro--trichloro-methylbenzhydrol. For the structure see Figure 93.1. Dicofol (BSI, ISO) is also called Kelthane (JMAF). Proprietary names include Acarin, Decofol, Hifol, Kelthane, and Mitigan. Dicofol has the empirical formula C14H9Cl5O and a molecular weight of 370.50. The pure substance is colorless and melts at 78.5– 70.5°C. It is soluble in most organic solvents but practically

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insoluble in water. Dicofol was introduced as a commercial chemical in 1955. Like chlorobenzilate, dicofol has been used mainly as a miticide for citrus fruits, nuts, cotton, and beans. It still appears to be used in some countries. The technical product is a brown viscous oil. The active compounds are 80% 1,1-bis(4-chlorophenyl)-2,2,2-trichloroethanol and 20% 1-(2-chlorophenyl)-1-(4-chloropenyl)-2,2,2-trichloroethanol (the o,p-isomer). The other major impurity detected is 1,1,1,2-tetrachloro-2,2-bis(4-chlorophenyl)ethane (Baum et al., 1976). Dicofol can be produced as water-dispersable powders, as emulsions, and in a dust.

93.7.2  Toxicity to Laboratory Animals In rats and rabbits, the acute oral LD50 for technical grade dicofol seems to range from 575 to 2000 mg/kg (Bendyke et al., 1970; Brown et al., 1969; Smith et al., 1959). Dogs seem to be much less sensitive (Smith et al., 1959). Rats fed dicofol for up to 2 years showed no effects on survival at levels below 1000 ppm but growth was impaired (Smith et al., 1959). The maximum tolerable dose for mice in a subchronic study was 500 ppm (Sato et al., 1987). Dogs fed 300 ppm showed no effect after 1 year, but some deaths occurred at 900 ppm. Dicofol seems to be metabolized in rats to 4,4-dichloro­ benzophenone, which is stored in fat and muscle as well as being excreted in the feces (Brown et al., 1969). DDE was also found, but there is doubt as to whether this was due to metabolism of dicofol or to contamination of the technical product employed. Water-soluble metabolites have been detected in the urine of mice give radiolabeled dicofol. Nearly 50% of the administered doses was excreted in the urine within 24 h. Part may be glucuronides of 4,4-dichlorobenzhydrol (Tabata et al., 1979). Brown and Casida showed that in vivo mice convert dicofol to dichlorobenzophenone and dichlorobenzhydrol and that DDE originates from the impurity -Cl-DDT (Brown and Casida, 1987). There is little published work on the specific toxic effects of dicofol in experimental animals. Some small adverse effects associated with reproduction in rats and mice have been reported (Jadaramkunti and Kaliwal, 2002; Trifonova and Gladenko, 1980) including inhibiting effects on implantation (Jadaramkunti and Kaliwal, 2001). In a comparative study, 98% dicofol, the technical product Kelthane, and DDT were given to male rats in equi­ molar amounts. Dicofol produced dosage-related increases in microsomal protein, cytochrome P450, and the specific activities of cytochrome reductase, ethoxycoumarin o-deethylase, aminopyrine N-demethylase, and glutathione S-transferase at a potency equivalent to that of Kelthane, DDT, or phenobarbital (Narloch et al., 1987). Some evidence has been obtained for its hepatocarcinogenicity in male B6C3F1 mice but not in rats (NCI, 1978d).

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93.7.3  Toxicity to Humans Only one case of possible human poisoning by dicofol seems to have been reported, and this was in combination with trichlorfon (Zolotnikova and Somov, 1978). Greenhouse workers reportedly suffered frequently from allergic dermatitis. A detailed study of the protection of workers in Florida citrus groves from contamination by dicofol has been reported (Nigg et al., 1986). Dicofol has cytokinetic and cytogenetic effects on human lymphoid cells in vitro (Sobti et al., 1983).

93.8.  Acetofenate 93.8.1  Identity, Properties and Uses Acetofenate is 2,2,2-trichloro-(3,4-dichlorophenyl) ethanol acetate or 2,2,2-trichloro-1(3,4-dichlorophenyl) ethyl acetate. For the structure see Figure 93.5. It is also called plifenate and apparently benzethazet or 7504. The CAS No. is 51366-277. Plifenate has a CAS No. 21757-82-4. The empirical formula is C10H7C15O2 and it has a molecular weight of 336.42. Although soluble in some solvents it has a water solubility of 40 mg/l. It is apparently of current use in China and nearby regions for mosquito control and flies, both indoors and outdoors, especially where resistance to DDT or lindane has developed or insecticides have been banned. The insecticide can be mixed with other pesticides in powders or solutions or incorporated into fibers having insect repellent properties.

93.8.2  Toxicity Overall there appears to be little easily accessible information on acetofenate uses and its metabolism and toxicity. Although not a double phenyl analogue of DDT, it does have many structural similarities (Figure 93.5) and is reportedly an efficient substitute. Although thought to be rapidly metabolized (Fytizas and Ioannou, 1982), compared with DDT there does not appear to be information on its bioaccumulation (Zhao et al., 2009). Oral administration of the insecticide to rats at 250 mg/kg gave acetofenate in adipose tissue at a maximum 24 h afterward with no accumulative effect if the dose was split over 15 days (Fytizas and Ioannou, 1982). No toxic effects were observed. In other studies with mice, oral LD50 values were established of 1164 mg/kg and 3116 mg/kg for male and female mice, respectively. Signs of dysfunction of the central nervous system and irritation of the gastrointestinal tract were observed. It was believed to be weakly accumulative. In a study of diets containing acetofenate for 10 months, enlarged liver was observed at the highest dose (500 ppm) (Wu et al., 1991). Recently, enantiomers of acetofenate (Figure 93.5) were separated and shown to affect 4-day zebrafish larvae differently in developmental toxicities such as yolk sac and pericardial edema, but not in acute toxicities (Xu et al., 2008). Quantitative RT-PCR of estrogen receptor  mRNA

CCl3 *

O

O CH3

Cl Cl Figure 93.5  Structure of acetofenate. The carbon atom at * is a chiral center with optical enantiomers.

gene expression showed that S-()-acetofenate was more potent than the R-() enantiomer, possibly associated with the developmental effects. The mixture of enantiomers of acetofenate was used to study the immunotoxicity toward a mouse macrophage cell line (Zhao et al., 2009). The insecticide caused apoptosis apparently associated with generation of reactive oxygen species and DNA damage signaling via the p53 pathway and decline in the Bcl-2/Bax ratio. As with the effects on zebra fish development, the S-() enantiomer was a more potent immunotoxin than the R-() isomer.

Conclusion Although organochlorine insecticides, pre-eminently DDT, have either been banned worldwide or have very restricted use, their toxicology is still of some interest and provides important lessons in pesticide toxicology. Because of the indiscriminate use of DDT in crop protection for a few decades, it persists widely in the biosphere and hence concerns for the implication to human health continue. In fact, the preceding sections show that a great deal is known about the storage, metabolism and toxicity of DDT and its analogues as far as humans and other mammals are concerned. Although effects on widelife due to accumulation up the food chain are well recorded, people seem to be remarkably resistant to any toxic effects of DDT even at very high exposure levels. Recent hypotheses about subtle effects on endocrine disruption, including some cancers, following chronic or neonatal exposures to these chemicals at low levels, although important, are still far from substantiated. These considerations have to be taken in the context of retaining DDT, in a more rational manner than in the past, as a potential weapon in fighting vector-transmitted diseases that may emerge in situations of resistance to current pesticides or in new areas of infection due to climate change. The experience and knowledge already gained will be also be very important in assessing the potential effects of emerging insecticides as exemplified here by acetofenate.

References Abou-Donia, M. B., and Menzel, D. B. (1968). The metabolism in vivo of 1,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane (DDT), 1,1-dichloro2,2-bis(p-chlorophenyl)ethane (DDD) and 1,1-dichloro-2,2-bis

Chapter | 93  Toxicology of DDT and Some Analogues

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Dodge, J. A., Glasebrook, A. L., Magee, D. E., Phillips, D. L., Sato, M., Short, L. L., and Bryant, H. U. (1996). Environmental estrogens: effects on cholesterol lowering and bone in the ovariectomized rat. J. Steroid Biochem. Mol. Biol. 59, 155–161. Domenjoz, R. (1944). Experimental investigation with a new insecticide (Neocide Geigy): A contribution to the theory of action of contact poison. Schwiz. Med. Wochenschr. 74, 952–958. Domenjoz, R. (1946). On the biological action of DDT-derivative. Arch. Int. Pharmacodyn. 73, 155–161. Draize, H., Nelson, A. A., and Calvery, H. O. (1944). The percutaneous absorption of DDT (2,2-bis-(p-chlorophenyl)-1,1,1-trichloroethane) in laboratory animals. J. Pharmacol. Exper. Ther. 82, 159–166. Dubois, J. M., and Bergman, C. (1977). Asymmetrical currents and sodium currents in Ranvier nodes exposed to DDT. Nature 266, 741–742. Durham, W. F., Ortega, P., and Hayes, J. Jr. (1963). The effect of various dietary levels of DDT on liver function cell morphology, and DDT storage in the rhesus monkey. Arch. Int. Pharmacodyn. Ther. 141, 111–129. Dutta, R., Mondal, A. M., Arora, V., Nag, T. C., and Das, N. (2008). Immunomodulatory effect of DDT (bis[4-chlorophenyl]-1,1,1-trichloroethane) on complement system and macrophages. Toxicology 252, 78–85. Epstein, S. S., and Shafner, H. (1968). Chemical mutagens in the human environment. Nature 219, 385–387. Eriksson, P., Falkeborn, Y., Nordberg, A., and Slanina, P. (1984). Effects of DDT on muscarine- and nicotine-like binding sites in CNS of immature and adult mice. Toxicol. Lett. 22, 329–334. Eroschenko, V. P., (1991). Ultrastructure of vagina and uterus in young mice after methoxychlor exposure. Reprod. Toxicol. 5, 427–435. Eroschenko, V. P., Abuelatta, A. A., and Grober, M. S. (1995). Neonatal exposures to technical methoxychlor alters ovaries in adult mice. Reprod. Toxicol. 9, 379–387. Eroschenko, V. P. and Cooke, P. S. (1990). Morphological and biochemical alterations in reproductive tracts of neonatal female mice treated with the pesticide methoxychlor. Biol. Reprod. 42, 573–583. Esaac, E. G., and Matsumura, F. (1980). Mechanisms of reductive dechlorination of DDT  sup>1  /sup> by rat liver microsomes. Pest. Biochem. Physiol. 13, 81–93. Eskenasy, J. J., (1972). Status epilepticus by dichlorodiphenyltrichlorethane and hexachlorcyclohexane poisoning. Rev. Roum. Neurol. 9, 435–442. Eyzaguirre, C., and Lilienthal, J. L. Jr. (1949). Veratrinic effects of pentamethylenetetrazol and 2,2-bis (p-chlorophenyl) 1,1,1 trichloroethane on mammalian neuromuscular function. Proc. Soc. Exp. Biol. Med. 70, 272–275. Falck, F. Jr., Ricci, A. Jr., Wolff, M. S., Godbold, J., and Deckers, P. (1992). Pesticides and polychlorinated biphenyl residues in human breast lipids and their relation to breast cancer. Arch. Environ. Health 47, 143–146. Farkas, I., Desi, I., and Kemeny, T. (1968). The effect of DDT in the diet on the resting and loading electrocorticogram record. Toxicol. Appl. Pharmacol. 12, 518–525. Fawcett, S. C., Bunyan, P. J., Huson, L. W., King, L. J., and Stanley, P. I. (1981). Excretion of radioactivity following the intraperitoneal administration of 14C-DDT, 14C-DDD, 14C-DDE and 14C-DDMU to the rat and Japanese quail. Bull. Environ. Contam. Toxicol. 27, 386–392. Fawcett, S. C., King, L. J., Bunyan, P. J., and Stanley, P. I. (1987). The metabolism of 14C-DDT, 14C-DDD, 14C-DDE and 14C-DDMU in rats and Japanese quail. Xenobiotica 17, 525–538.

Hayes’ Handbook of Pesticide Toxicology

Fei, X., Chung, H., and Taylor, H. S. (2005). Methoxychlor disrupts uterine Hoxa10 gene expression. Endocrinology 146, 3445– 3451. Feil, V. J., Lamoureux, C. J., Styrvoky, E., Zaylskie, R. G., Thacker, E. J., and Holman, G. M. (1973). Metabolism of o,p-DDT in rats. J. Agric. Food Chem. 21, 1072–1078. Feil, V. J., Lamoureux, C. H., and Zaylskie, R. G. (1975). Metabolism of o,p-DDT in chickens. J. Agric. Food Chem. 23, 382–388. Fennah, R. G. (1945). Preliminary tests with DDT against insect pests of food crops in the Lesser Antilles. Tropic. Agric. 22, 222–226. Finnegan, J. K., Hennigar, G. R., Smith, R. B. Jr., Larson, P. S., and Haag, H. B. (1955). Acute and chronic toxicity studies on 2,2-bis-(p-ethylphenyl)-1,1-dichloroethane (perthane). Arch. Int. Pharmacodyn. Ther. 103, 404–418. Fisher, S. G., and Fisher, R. I. (2004). The epidemiology of nonHodgkin’s lymphoma. Oncogene 23, 6524–6534. Fitzhugh, O. G. (1948). Use of DDT insecticides on food products. Ind. Eng. Chem. 40, 704–705. Fitzhugh, O. G. (1970). A summary of carcinogenic study of DDT in mice. In “FAO/WHO 1969 Evaluation of Some Pesticides Residues in Foods” FAO pp. 61–64. Fitzhugh, O. G., and Nelson, A. A. (1947). The chronic oral toxicity of DDT (2,2,-bis-p-chlorophenyl-1,1,1-trichoroethane). J. Pharmacol. Exper. Therap. 89, 18–30. Flodstrom, S., Hemming, H., Warngard, L., and Ahlborg, U. G. (1990). Promotion of altered hepatic foci development in rat liver, cytochrome P450 enzyme induction and inhibition of cell-cell communication by DDT and some structurally related organohalogen pesticides. Carcinogenesis 11, 1413–1417. Fonseca, M. I., Aguilar, J. S., Lopez, C., Garcia Fernandez, J. C., and De Robertis, E. (1986). Regional effect of organochlorine insecticides on cholinergic muscarinic receptors of rat brain. Toxicol. Appl. Pharmacol. 84, 192–195. Foster, T. S. (1973). Evaluation of the possible estrogenic activity of methoxychlor in the chicken by means of feeding trials. Bull. Environ. Contam. Toxicol. 9, 234–242. Francone, M. P., Mariani, F. H., and Demare, C. (1952). Clinical aspects of DDT poisoning. Rev. Asoc. Med. Argent. 66, 56–59. French, M. C., and Jefferies, D. J. (1969). Degradation and disappearance of ortho, para isomer of technical DDT in living and dead avian tissues. Science 165, 914–916. Fries, G. F., Marrow, G. S. Jr., Lester, J. W., and Gordon, C. H. (1971). Effect of microsomal enzyme inducing drugs on DDT and dieldrin elimination from cows. J. Dairy Sci. 54, 364–368. Frigo, D. E., Burow, M. E., Mitchell, K. A., Chiang, T. C., and McLachlan, J. A. (2002). DDT and its metabolites alter gene expression in human uterine cell lines through estrogen receptor-independent mechanisms. Environ. Health Perspect. 110, 1239–1245. Frigo, D. E., Tang, Y., Beckman, B. S., Scandurro, A. B., Alam, J., Burow, M. E., and McLachlan, J. A. (2004). Mechanism of AP-1mediated gene expression by select organochlorines through the p38 MAPK pathway. Carcinogenesis 25, 249–261. Fryzek, J. P., Garabrant, D. H., Harlow, S. D., Severson, R. K., Gillespie, B. W., Schenk, M., and Schottenfeld, D. (1997). A case-control study of self-reported exposures to pesticides and pancreas cancer in southeastern Michigan. Int. J. Cancer 72, 62–67. Fukushima, D. K., Bradlow, H. L., and Hellman, L. (1971). Effects of o,p-DDD on cortisol and 6-beta-hydroxycortisol secretion and metabolism in man. J. Clin. Endocrinol. Metab. 32, 192–200.

Chapter | 93  Toxicology of DDT and Some Analogues

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Goldman, M. (1981). The effect of a single dose of DDT on thyroid function in male rats. Arch. Int. Pharmacodyn. Ther. 252, 327–334. Goldman, J. M., Cooper, R. L., Rehnberg, G. L., Hein, J. F., McElroy, W. K., and Gray, L. E. Jr. (1986). Effects of low subchronic doses of methoxychlor on the rat hypothalamic-pituitary reproductive axis. Toxicol. Appl. Pharmacol. 86, 474–483. Gordienko, V. M., Bogomolets, Y. O., and Kozyritskii, V. G. (1972). Electron microscopic study of the dog thyroid gland following differerent doses of o,p-DDD. Tsitol. Genet. 6, 392–394. Gordienko, V. M., and Kozyritskii, V. G. (1970). Effect of ortho-, para-, prime-dichlordiphenyldichlorethane on the ultrastructure of cells of the anterior pituitary gland in dogs. Arkh. Anat. Gistol. Embriol. 58, 49–57. Gordienko, V. M., and Kozyritskii, V. G. (1973). Change in the ultrastructure of the dog adrenal cortex following short-term and prolonged administration of o, p-DDD. Arkh. Anat. Gistol. Embriol. 65, 90–95. Gordon, H. T., and Welsh, J. H. (1948). The role of ions in axon surface reactions to toxic organic compounds. J. Cell Physiol. 31, 395–419. Graillot, C., Gak, J. C., Lancret, C., and Truhaut, R. (1975). Research on the modalities and mechanisms of toxic action of organochlorine insecticides. II. Long term toxicity effects of DDT in hamsters. Eur. J. Toxicol. Environ. Hygiene – J. Eur. Toxicol. 8, 353–359. Grant, E. L., Mitchell, R. H., West, P. R., Mazuch, L., and AshwoodSmith, M. J. (1976). Mutagenicity and putrative carcinogenicity tests of several polycyclic aromatic compounds associated with impurities of the insecticide methoxychlor. Mutat. Res. 40, 225–228. Gray, L. E. Jr., Ostby, J. S., Ferrell, J. M., Sigmon, E. R., and Goldman, J. M. (1988). Methoxychlor induces estrogen-like alterations of behavior and the reproductive tract in the female rat and hamster: effects on sex behavior, running wheel activity, and uterine morphology. Toxicol. Appl. Pharmacol. 96, 525–540. Gray, L. E. Jr., Ostby, J., Ferrell, J., Rehnberg, G., Linder, R., Cooper, R., Goldman, J., Slott, V., and Laskey, J. (1989). A dose-response analysis of methoxychlor-induced alterations of reproductive development and function in the rat. Fundam. Appl. Toxicol. 12, 92–108. Griffith, F. D. Jr., and Blanke, R. V. (1975). Blood organochlorine pesticide levels in Virginia residents. Pestic. Monit. J. 8, 219–224. Guan, X., and Ruch, R. J. (1996). Gap junction endocytosis and lysosomal degradation of connexin 43-P2 in WB-F344 rat liver epithelial cells treated with DDT and lindane. Carcinogenesis 17, 1791–1798. Gupta, R. K., Schuh, R. A., Fiskum, G., and Flaws, J. A. (2006). Methoxychlor causes mitochondrial dysfunction and oxidative damage in the mouse ovary. Toxicol. Appl. Pharmacol. 216, 436–445. Gutierrez, M. L., and Crooke, S. T. (1980). Mitotane (o,p-DDD). Cancer Treat. Rev. 7, 49–55. Guttes, S., Failing, K., Neumann, K., Kleinstein, J., Georgii, S., and Brunn, H. (1998). Chlororganic pesticides and polychlorinated biphenyls in breast tissue of women with benign and malignant breast disease. Arch. Environ. Contam. Toxicol. 35, 140–147. Haag, H. B., Finnegan, J. K. et al (1948). Comparative chronic toxicity for warm-blooded animals of 2,2-bis-(p-chlorophenyl)-1, 1, 1-trichloroethane and 2,2-bis-(p-chlorophenyl)-1, 1-dichloroethane. Ind. Med. Surg. 17, 477–484. Hajjar, R. A., Hickey, R. C., and Samaan, N. A. (1975). Adrenal cortical carcinoma. A study of 32 patients. Cancer 35, 549–554. Hall, D. L., Payne, L. A., Putnam, J. M., and Huet-Hudson, Y. M. (1997). Effect of methoxychlor on implantation and embryo development in the mouse. Reprod. Toxicol. 11, 703–708. Halmi, K. A., and Lascari, A. D. (1971). Conversion of virilization to feminization in a young girl with adrenal cortical carcinoma. Cancer 27, 931–935.

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Hamid, J., Sayeed, A., and McFarlane, H. (1974). The effect of 1-(o-chlorophenyl)-1-(p-chlorophenyl)-2,2-dichloroethane (o,p-DDD) on the immune response in malnutrition. Br. J. Exp. Pathol. 55, 94–100. Harell, M., Shea, J. J., and Emmett, J. R. (1978). Bilateral sudden deafness following combined insecticide poisoning. Laryngoscope 88, 1348–1351. Harris, S. J., Cecil, H. C., and Bitman, J. (1974). Effect of several dietary levels of technical methoxychlor on reproduction in rats. J. Agric. Food Chem. 22, 969–973. Harrison, R. D. (1975). Comparative toxicity of some selected pesticides in neonatal and adult rats. Toxicol. Appl. Pharmacol. 32, 443–446. Hart, M. M., and Straw, J. A. (1971a). Effect of 1-(o-chlorophenyl)-1(p-chlorophenyl)-2,2-dichloroethane on adrenocorticotropic hormone-induced steroidogenesis in various preparations in vitro of dog adrenal cortex. Biochem. Pharmacol. 20, 1679–1688. Hart, M. H., and Straw, J. A. (1971b). Effect of 1-(o-chlorophenyl)-1-(pchlorophenyl)-2, 2-dichloroethane in vivo on baseline and adrenocorticotropic hormone-induced steroid production in dog adrenal slices. Biochem. Pharmacol. 20, 1689–1691. Hart, M. M., and Straw, J. A. (1971c). Effects of 1-(o-chlorophenyl)-1(p-chloro-phenyl)-2,2-dichloroethane and puromycin on adrenocorticotropic hormone-induced steroidogenesis and on amino acid incorporation in slices of dog adrenal cortex. Biochem. Pharmacol. 20, 257–263. Hart, M. M., and Straw, J. A. (1971d). Studies on the site of action of o,p-DDD in the dog adrenal cortex. 1. Inhibition of ACTH-mediated pregnenolone synthesis. Steroids 17, 559–574. Hart, M. M., Reagan, R. L., and Adamson, R. H. (1973). The effect of isomers of DDD on the ACTH-induced steroid output, histology and ultrastructure of the dog adrenal cortex. Toxicol. Appl. Pharmacol. 24, 101–113. Hartwig, W., Massalski, W., Kasperlik-Zaluska, A., Migdalska, B., Szamatowicz, M., and Jakowicki, J. (1968). Hormonally acitve cancer of the adrenal cortex treated with o,p-DDD. Endokrynol. Pol. 19, 187–200. Hassall, K. A. (1971). Reductive dechlorination of DDT: The effect of some physical and chemical agents on DDE production by pigeon liver preparations. Pestic. Biochem. Physiol. 1, 259–266. Hayes, W. J., and Laws, E. R. (1991). “Handbook of Pesticide Toxicology”. Academic Press, San Diego. Hayes, W. J. Jr. (1959). Pharmacology and toxicology of DDT. In “The Insecticide Dichlorodiphenyl-Trichloroethane and its Significance” (P. Muller, ed.). Birkhauser, Basel. Hayes, W. J. Jr. (1965). Review of the metabolism of chlorinated hydrocarbon insecticides especially in mammals. Ann. Rev. Pharmacol. 10, 27–52. Hayes, W. J. Jr. (1975). “Toxicology of Pesticides”. Williams and Wilkins, Baltimore. Hayes, W. J. Jr. (1976). Dosage relationships associated with DDT in milk. Toxicol. Appl. Pharmacol. 38, 19–28. Hayes, W. J. Jr. (1982). “Pesticides Studied in Man”. Williams and Wilkins, Baltimore. Hayes, W. J. Jr. (1991). Introduction. In “Handbook of Pesticide Toxicology” (W. J. Hayes and E. R. Laws, eds.), pp. 1–37. Academic Press, San Digeo. Hayes, W. J. Jr., Dale, W. E., and Pirkle, C. I. (1971). Evidence of safety of long-term, high, oral doses of DDT for man. Arch. Environ. Health 22, 119–135.

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1,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane (DDT) to 1,1-dichloro2,2-bis(p-chlorophenyl)ethane (DDD). Biochem. Pharmacol. 35, 1805–1807. Keplinger, M. L., Deichmann, W. B., and Sala, F. (1970). Effects of combinations of pesticides on reproduction in mice. In “Pesticides Symposia” (W. B. Deichmann, J. L. Radomski, and R. A. Penalver, eds.), pp. 125–138. Halos and Associates, Miami, FL. Key, T., and Reeves, G. (1994). Organochlorines in the environment and breast cancer. Br. Med. J. 308, 1520–1521. Khaikina, B. I., and Shilina, V. F. (1971). Effect of some chlororganic pesticides on serotonin metabolism. Farmakol. Toksikol. 34, 357–359. Khairy, M. (1959). Changes in behaviour associated with a nervous system poison (DDT). Q. J. Exp. Physiol. 11, 91–94. Khera, K. S., Whalen, C., and Trivett, G. (1978). Teratogenicity studies on linuron, malathion, and methoxychlor in rats. Toxicol. Appl. Pharmacol. 45, 435–444. Kishimoto, D., and Kurihara, N. (1996). Effects of cytochrome P450 antibodies on the oxidative demethylation of methoxychlor catalyzed by rat liver microsomal cytochrome P450 isozymes: Isozyme specificity and alteration of enantiotopic selectivity. Pestic. Biochem. Physiol. 56, 44–52. Kishimoto, D., Oku, A., and Kurihara, N. (1995). Enantiotopic selectivity of cytochrome P450-catalyzed oxidative demethylation of methoxychlor – alteration of selectivity depending on isozymes and substrate concentrations. Pestic. Biochem. Physiol. 51, 12–19. Kitamura, S., Shimizu, Y., Shiraga, Y., Yoshida, M., Sugihara, K., and Ohta, S. (2002). Reductive metabolism of p,p-DDT and o,p-DDT by rat liver cytochrome. Drug Metab. Dispos. 30, 113–118. Klein, A. K., Laug, E. P., Datta, P. R., and Mendel, J. L. (1965). Evidence for the conversion of o,p-DDT (1,1,1-trichloro-2-o-chlorophenyl-2-pchlorophenylethane) to p,p-DDtT (1,1,1-trichloro-2,2-bis(p-chloro­ phenyl)ethane in rats. J. Am. Chem. Soc. 87, 2520–2522. Klotz, H. P., Thibaut, E., and Russo, F. (1971). Low doses of o,p-DDD in anovulatory spanomenorrheas of hypertrichosic patients. Ann. Endocrinol. (Paris) 32, 763–767. Knowles, C. O., and Ahmad, S. (1971). Comparative metabolism of chlorobenzilate, chloropropylate, and bromopropylate acaricides by rat hepatic enzymes. Can. J. Physiol. Pharmacol. 49, 590–597. Kolmodin, B., Azarnoff, D. L., and Sjoqvist, F. (1969). Effect of environmental factors on drug metabolism: decreased plasma half-life of antipyrine in workers exposed to chlorinated hydrocarbon insecticides. Clin. Pharmacol. Ther. 10, 638–642. Komissarenko, V. P. (1971). Effect of o,p-DDT (clodithane) on secretion and metabolism of corticosteroids in chickens. Fiziol. Zh. Kiev. 17, 435–441. Komissarenko, V. P., Gordienko, V. M., and Reznikov, A. G. (1972). Restorative processes in the adrenal cortex of dogs after administration of o,p-DDD. Probl. Endokrinol. (Mosk). 18, 74–81. Komissarenko, V. P., Chelnakova, I. S., and Mikosha, A. S. (1978). Activity of glutathione reductase in the adrenal glands and the liver of dogs after administration o,p-DDD, perthane and ACTH. Probl. Endokrinol. (Mosk). 24, 95–98. Komissarenko, V. P., Reznikov, A. G., Gordienko, V. M., and Zak, K. P. (1968). Effect of o,p-DDD on the morphology and function of adrenal cortex in dogs. Endocrinol. Exp. 2, 21–28. Korpachev, V. V. (1972a). Accumulation and elimination of o,p-DDD in organs and tissues of guinea pigs and dogs. Fiziol. Z. 18, 585–590. Korpachev, V. V. (1972b). Dependence of o,p-DDD absorption on dose and drug form. Farm. Zh. 27, 64–66.

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Chapter | 93  Toxicology of DDT and Some Analogues

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Chapter | 93  Toxicology of DDT and Some Analogues

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Zachow, R., and Uzumcu, M. (2006). The methoxychlor metabolite, 2,2-bis-(p-hydroxyphenyl)-1,1,1-trichloroethane, inhibits steroidogenesis in rat ovarian granulosa cells in vitro. Reprod. Toxicol. 22, 659–665. Zaidi, S. S. (1987). Possible role of rat liver NADPH cytochrome P-450 reductase in the detoxication of DDT to DDD. Bull. Environ. Contam. Toxicol. 39, 327–333. Zein-el-Dine (1946). The insecticide DDT. J. Egypt. Med. Assoc. 29, 38–54. Zhao, M., Zhang, Y., Wang, C., Fu, Z., Liu, W., and Gan, J. (2009). Induction of macrophage apoptosis by an organochlorine insecticide acetofenate. Chem. Res. Toxicol. 22, 504–51. Zhou, L. X., Dehal, S. S., Kupfer, D., Morrell, S., McKenzie, B. A., Eccleston, E. D. Jr., and Holtzman, J. L. (1995). Cytochrome P450 catalyzed covalent binding of methoxychlor to rat hepatic, microsomal iodothyronine 5-monodeiodinase, type I: does exposure to methoxychlor disrupt thyroid hormone metabolism? Arch. Biochem. Biophys. 322, 390–394. Ziem, G. (1982). Aplastic anaemia after methoxychlor exposure. Lancet 2, 1349. Zolotnikova, G. P., and Somov, B. A. (1978). Role of pesticides in occurrence of occupational dermatites in workers of hothouses. Vestn. Dermatol. Venerol., 76–79. Zumoff, B. (1979). The hypouricemic effect of o,p-DDD. Am. J. Med. Sci. 278, 145–147.

Chapter 94

Boric Acid and Inorganic Borate Pesticides Craig E. Bernard, Michael C. Harrass and Mark J. Manning Rio Tinto Borax, Boron, California

94.1  Introduction 94.1.1  Background As one of the 109 elements that make up the planet, it is not surprising that boron is everywhere – in soil and water, plants and animals – in trace amounts. Although scientists refer to levels of “boron,” it is important to note that the element boron does not exist by itself in nature. Rather, boron combines with oxygen and other elements to form boric acid, or inorganic borate salts called borates. Boric acid and borates are water-soluble white powder substances with low acute oral and dermal toxicity that are widely used for a variety of industrial purposes including manufacture of glass, insulation fiberglass (IFG), porcelain enamel, ceramic glazes, and metal alloys. They are also used as fire retardants, laundry additives, fertilizers (boron is an essential element for plants), herbicides (at high concentrations, boron is toxic to certain plant species), and insecticides (Woods, 1994). Elemental boron itself has only limited industrial applications. Registered pesticide applications have been around since the 1940s

with use as acaricides, algaecides, fungicides, herbicides, and insecticides. A number of detailed hazard assessments and reviews of the toxicology of borates have been published (ATSDR, 1993; Culver and Hubbard, 1996; Culver et al., 1994b; ECETOC, 1995; EFSA, 2004; European Commission, 1996; Hubbard, 1998; Hubbard and Sullivan, 1996; IPCS, 1998; Murray, 1995; WHO, 1998; Moore et al., 1997; UK EVM, 2003; U.S. Food and Nutrition Board, 2001; U.S. EPA, 1993, 2004, 2006a,b). This includes ­ evaluations from various review bodies and regulatory agencies which include guidance on acceptable references doses as described in Table 94.1 below.

94.1.2  Chapter Coverage This chapter focuses on boric acid and inorganic borates, primarily borax (e.g., disodium tetraborate decahydrate), disodium octaborate tetrahydrate, and their derivatives, since the major pesticide uses involve these materials. This presentation is not intended to be a comprehensive treatise on

Table 94.1  Chemical-Specific Adjustment Factors in Various Risk Assessments of Boron-Containing Compounds Uncertainty factor

IEHR (1995)

ECETOC (1995)

IPCS (1998)

WHO (1998)

NAS FNB (2000)

UK EFSA (2004)

US EPA IRIS (2004)

Interspecies

3.2

3

3.2

10

10

10

10

Intraspecies

10

10

7.9

6

3

6

6

Total

32

30

25

60

30

60

66

Reference dose (mg boron/day)

0.3

0.32

0.4

0.16

0.3

0.16

0.2 (rounded from 0.16)

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the chemical or toxicological nature of boron compounds but rather provide a working knowledge of the relevant acute and chronic toxicological studies of boric acid and inorganic borate salts. Most of the simple inorganic borates exist predominantly as undissociated boric acid in dilute aqueous solution at physiological pH, leading to the conclusion that the main species in the plasma of mammals is undissociated boric acid. Since other borates dissociate to form boric acid in aqueous solutions, they too can be considered to exist as undissociated boric acid under physiological conditions. For example disodium octaborate tetrahydrate, or DOT, can be thought of as a solid, amorphous mixture of boric acid and disodium tetraborate decahydrate. For this reason, the majority of toxicological studies of borates have involved either boric acid or disodium tetraborate decahydrate (e.g., borax). Both acute and longer-term studies have been conducted using these two substances. For boric oxide and other sodium borates such as disodium tetraborate pentahydrate, disodium tetraborate anhydrous and disodium octaborate tetrahydrate, only acute toxicity studies have been carried out. For comparative purposes, dose levels of borates are usually expressed in terms of boron (B) equivalents based on the fraction of boron content on a molecular weight basis. Conversion factors are given in Table 94.2. These conversion factors are important as some studies express dose in terms of B, whereas other studies express the dose in units of parent compound (i.e., boric acid or disodium tetraborate decahydrate). The B equivalents used are a generic designation rather than a designation of the element boron.

94.1.3  Product Uses Boric acid was first registered in the United States for pesticide use in 1948. As of June 2009, there were 64 registered products containing boric acid, 17 registered products containing disodium tetraborate decahydrate, 37 registered products containing disodium octaborate tetrahydrate, and four registered products containing zinc borate. Many formulations are marketed, including liquids, soluble

and emulsifiable concentrates, granulars, powders, dusts, ­pellets, tablets, solids, paste, baits, and crystalline rods. Uses include insecticides, fungicides, and algaecides, with very little herbicide usage remaining. The major borate pesticide usage includes control of cockroaches, ants, wooddestroying insects (including termites), and fungi. This topic has been recently reviewed (Lloyd, 1998). Borates are used for stump treatment (control of Fomes annosus) and algae control in swimming pools. They are also used for flea control in carpets, although this use has been controversial due to some high application rates reported and the resultant concern for exposures to children and pets. Boric acid and inorganic borate salt compounds are frequently used for control of insects such as ants or roaches by application in nonagricultural food and feed areas. Boric acid and sodium borate salts have herbicide qualities causing desiccation, fungicidal activity through fungi growth inhibition by preventing the production of conidia or asexual spores, and insecticidal properties by acting as a stomach poison against ants, cockroaches, silverfish, and termites. Use sites include animal housing, turf, transportation and storage facilities, medical/veterinary institutions, uncultivated agricultural/nonagricultural areas, refuse/solid waste sites, swimming pool (algae control), paved areas, and aquatic structures. Borates have also been used as broad-spectrum wood preservatives for over 70 years due to their excellent efficacy against various wood-destroying organisms (fungi, wood boring beetles, and termites). Boric acid/sodium borate salts are used as active and inert ingredients in pesticide formulations. Residential use products are most commonly applied indoors as a broadcast spray (e.g., decks and carpets) and crack and crevice treatments. Borates are typically applied as wood preservatives through pressure treatment or spray application of the lumber and/or plywood with aqueous solutions, either with borates alone or where they are used as part of more complex formulations. Boric acid/sodium borate salts are added to suppress algae growth in swimming pools and as a buffer and chlorine extender. There is a variety of use patterns and formulations (e.g., liquids, dusts, and granulars) for boric acid/sodium borate salts.

Table 94.2  Conversion Factors to Boron Equivalents Boron species

Formula

Conversion factor for equivalent dose of B

Boric acid

H3BO3

0.175

Disodium tetraborate decahydrate (Borax)

Na2B4O7 ·10H2O

0.113

Disodium octaborate tetrahydrate (DOT)

Na2B8O13 · 4H2O

0.21

Zinc borate

2ZnO•3B2O3•3H2O

0.149

Chapter | 94  Boric Acid and Inorganic Borate Pesticides

Several types of borate wood preservatives are currently used to treat solid wood, engineered wood composites, and other interior building products like studs, plywood, joists, and rafters. These include disodium octaborate tetrahydrate and zinc borate. Zinc borate is prepared by reacting boric acid and zinc oxide. Zinc borate, which has low aqueous solubility compared to boric acid and sodium borate salts, is much more resistant to leaching, thereby making it potentially suitable for rigorous end-use exposures, such as unprotected, above-ground applications, where soluble borates are not normally used. Because of this reduced solubility, zinc borate cannot be used in pressure treating processes like boric acid/sodium salts and is typically incorporated as a powder into a wood composite or wood-plastic composite during the manufacturing process. Borates interfere with termites’ metabolic pathways when ingested through feeding or grooming, effectively killing them. Surviving termites avoid the protected wood products (Grace, 1997).

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20°C; pH, 8.5 (1% at 23°C); bulk density, 320–480 kg/m3. The chemical composition formula given above is an approximate composition of the solid material. The solid is amorphous and the exact structure is not known. It is the composition that gives the maximum water solubility of all the sodium borate species.

94.2.3  Zinc Borate Zinc borate is the common name (zinc borate hydrate 2335, dodecaboron, tetrazinc docosaoxide, heptahydrate) and 2ZnO•3B2O3•3H2O the chemical name. The CAS registry number is 138265-88-0. It appears as a white crystalline powder; molecular weight, 434.66; melting point, phase change at 650°C (1202°F); vapor pressure, negligible at 20°C; water solubility, less than 0.28% at 25°C; pH, 6.8–7.5 (aqueous solution) at 20°C; specific gravity, 2.7.

94.3  Exposure 94.2  Names and chem�������������� i������������� cal/physical properities 94.2.1  Boric Acid Boric acid is the common name (boracic acid, orthoboric acid) and B(OH)3 the chemical name. The CAS registry number is 10043-35-3. It appears as a white crystalline powder; molecular weight, 61.88; melting point, 170.9°C  0.2°C; vapor pressure, less than 10−4 torr at 20°C; water solubility, 4.72% at 20°C; octanol/water partition coefficient, 0.175; pH: 5.1 (1% solution) at 20°C; specific gravity, 1.51.

94.2.2  Sodium Borate Salts 94.2.2.1  Disodium Tetraborate Decahydrate Disodium tetraborate decahydrate (sodium tetraborate decahydrate, borax, borax decahydrate, borax 10-mole), Na2B4O7 •10H2O (CAS No. 1303-96-4). It appears as a white crystalline powder; molecular weight, 381.87; melting point, 62°C (it begins to dissolve in water of hydration); vapor pressure, less than 10−6 torr at 20°C; water solubility, 4.71% at 20°C; pH, 9.24 (1% solution) at 20°C; specific gravity, 1.71.

94.2.2.2  Disodium Octaborate Tetrahydrate Disodium octaborate tetrahydrate is the common name and Na2B8O13• 4H2O the chemical name. The CAS registry number is 12280-03-4. It appears as a white amorphous powder; molecular weight, 412.52; melting point, 815°C; vapor pressure, negligible at 20°C; water solubility, 9.5% at

94.3.1  Dietary Exposure The primary source of exposure for human populations is ingestion of boron from food (primarily fruits and vegetables) (Coughlin, 1998; Meacham and Hunt, 1998; Rainey and Nyquist, 1998; WHO, 1998). This is largely due to the essentiality of boron as a micronutrient in plants (Woods, 1994), which has been known since the 1920s. Thus, plantderived foods such as fruits, vegetables, and nuts contain significant amounts of boron. Foods with the highest concentrations of boron include avocado, peanut butter, peanuts, prune juice, grape juice, chocolate powder, wine, pecans, granola raisin cereal, and raisin bran cereal (Meacham and Hunt, 1998). In the United States, the greatest contribution of boron to the diet is from coffee, milk, apples, dried beans, and potatoes, which together account for 27% of dietary boron consumption (Rainey et al., 1999, 2002). The contribution from coffee and milk stems from their consumption in large quantities, not because they are high in B content.

94.3.2  Occupational Exposure Occupational exposure is included in the presentation below under Sections 94.8.1 and 94.8.2.

94.3.3  Residential Exposure The U.S. EPA (2006b) recently evaluated the human health risks associated with existing product registrations in the United States. Their assessment included postapplication exposure from broadcast carpet application, crackand-crevice treatments, pressure-treated wood, and deck treatment from spraying (in situ). A primary concern was

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exposure for children resulting from incidental ingestion of residues due to surface-to-hand-to-mouth transfer. The estimated average absorbed daily dosages for young children based on hand-to-mouth transfer from the treated surface for broadcast carpet application, crack-and-crevice treatments, pressure-treated wood, and deck treatment from spraying (in situ) were 0.05, 0.03, 0.015, and 0.05 mg/kg bw/day, respectively. Based on these estimated exposures, the agency determined that there was a reasonable certainty that no harm will result from aggregate nonoccupational exposure to the pesticide chemical risks for most of the scenarios assessed. Anticipated exposure from swallowing swimming pool water containing disodium tetraborate pentahydrate from products applied to both swimming pools and spas were also evaluated. The assessment included a survey of the range of maximum application rates from product labels. The rates reported ranged from 2 lb a.i. per 1000 gallons (240 mg/l) to 4.5 lb a.i. per 1000 gallons (540 mg/l pool water concentration). The calculated absorbed daily dose or ADD for a child aged 7–10 years, based on the U.S. EPA SWIMODEL (Version 2.0), ranged from 0.059 to 0.134 mg/kg bw/day. Krieger et al. (1996) evaluated exposure potential of disodium octaborate tetrahydrate for its use as an indoor flea control on carpets and furniture. Disodium octaborate tetrahydrate was applied to a nylon carpet using a powered rug brush at a rate of approximately 0.2 mg/cm2 carpet. Volunteers wore either bathing suits that provided 75% or more skin exposure or whole-body, cotton dosimeters consisting of socks, union suits, and gloves. The volunteers performed a 20-min set of Jazzercise routines, which was designed to represent a person’s day-long occasional contact with treated surfaces. The availability of boron was demonstrated by covering portions of the carpet with a cotton dosimeter and rolling it with a weighted roller. Additionally, the disodium octaborate tetrahydrate transferred was measured from the whole-body dosimeters. No evidence of contact transfer and dermal absorption was obtained. The mean daily boron levels (mg/g creatinine) from urine monitoring ranged from 1.17 to 1.33 for the group with exposed skin and 1.12 to 1.26 for those who wore dosimeters that prevented direct skin contact. Daily urine boron levels were not significantly different when compared using a two-sample t-test assuming equal variances (p  0.05). The study authors concluded that direct dermal contact with disodium octaborate tetrahydrate treated carpet at a nominal rate of 0.2 mg/cm2 did not produce any adverse effects or change urinary boron clearance. Inert uses such as laundry detergent and general ­purpose cleaners were also evaluated by the agency with calculated exposures for these residential uses being 16,000 or more below the toxicological endpoint of concern identified by the agency. The reader is referred to Human and Environmental Risk Assessment on ingredients of household

Hayes’ Handbook of Pesticide Toxicology

cleaning products or HERA (http://www.heraproject.com/ RiskAssessment.cfm) for additional estimates of exposure for boric acid and borate use in household cleaning products.

94.3.4  Environmental Exposure Terrestrial systems receive boron from atmospheric deposition, burning of wood and coal, and rock weathering. Sea salt aerosols are the largest source of atmospheric flux (1.44 Tg B/year; Park and Schlesinger, 2002). Weathering of rocks is a much smaller source (0.19 Tg B/year), about equal to coal combustion (0.20 Tg B/year) and less than burning biomass (0.33 Tg B/year). Global industrial use of boron is estimated at about 0.31 Tg B/year (Argust, 1998), with about half of that production being used in immobilized forms (e.g., glass). Of an estimated 2200 Tg B in surface soil, about 4.5 Tg B/year is taken in by plants globally, with most being returned to the soil as organic matter (Park and Schlesinger, 2002). Soil concentrations normally range from 2 to 100 ppm (Adriano et al., 1988). Soil concentrations reflect the underlying rock, with typical boron concentrations increasing from igneous rock, sedimentary rock, arid saline soils, and marine shale. Typical soil concentrations are reported as 10–33 ppm (Eisler, 1990; Lindsay, 1979). Freshwater concentrations range from nondetectable to several mg B/l, with the highest occurrences usually associated with geothermal inputs, such as the Firehole River in Yellowstone (Meyer et al., 1998). Sea water contains about 5 mg B/l. A survey of public water supplies in the United States found that the median boron concentration was 0.04 mg B/l with a maximum of 3.3 mg B/l and average of 0.17 mg B/l (Frey et al., 2004). This reflected 134 surface water sources and 273 groundwater sources. The U.S. EPA reported results from a survey of ground water sources used for drinking water and found that boron was detected (0.005 mg B/l) in about 82% of all ground waters, but exceeded 0.7 mg B/l in only about 4.3% of the samples, and exceeded 1.4 mg B/l in 1.7% of the ground waters used as public water supplies (U.S. EPA, 2006a). Wastewater containing anthropogenic boron has resulted in locally elevated boron concentrations. Boron is relatively mobile in the environment compared to other inorganic substances. Boric acid and borates are water-soluble and have relatively low binding to soil, as indicated by low Freundlich partitioning coefficients of about 1.6.

94.4  Biological importance 94.4.1  Essentiality in Plants Boron is a naturally occurring element that is essential to a variety of organisms. In plants, it is necessary for a variety of metabolic processes (e.g., nitrogen metabolism, nucleic acid metabolism, and membrane integrity and stability) and has

Chapter | 94  Boric Acid and Inorganic Borate Pesticides

been known to be an essential micronutrient for terrestrial plants for several decades (Butterwick et al., 1989; Eisler, 2000; Gupta. et al., 1985). Shorrocks (1997) documented the use of boron applications for 132 crops in over 80 countries, demonstrating the widespread nature of agricultural use of boron. Evidence exists that it is essential for nitrogen fixation in some species of algae (Smith and Dugger, 1981), fungi and bacteria (Fernandez et al., 1984; Saiki et al., 1993), some diatoms, and algae and macrophytes (Eisler, 2000). Required levels may vary, especially among plants, such that essential levels for one species may be toxic to another (Eisler, 1990). The concentration–response curve for boron is likely to be U-shaped for many species, with adverse effects observed at very high and very low concentrations, while no adverse effects are observed at the intermediate concentrations (Loewengart, 2001). Plant species vary in the concentrations associated with deficiency and toxicity. Monocotyledons (e.g., corn and grasses) require about onequarter as much boron as dicotyledons (e.g., tomatoes, carrots, clovers, beets) (Butterwick et al., 1989). The mobility of boron within the plant may help explain the observed deficiency and toxicity patterns. Boron is more mobile in plants that produce the simple sugars known as polyols (e.g., sorbitol and mannitol) than in species that do not produce polyols. In polyol-producing species, boron is translocated from one part of the plant to another and so may reach the meristem and affect growth. In the absence of polyols, boron is relatively immobile within the plant (Brown et al., 2002). A polyol-producing plant may be both more tolerant of boron deficiency and more sensitive to higher boron concentrations because of the mobility of boron within the plant. This is important in agricultural applications of boron, which may be applied as a soil treatment or as foliar spray. Agricultural application of boron depends on the plant and cultivar, as well as the local soil. Recommended application rates range from 0 to 4.5 kg B/ha (U.S. Borax, 2002), but typically are in the range of 1–2 kg B/ha (Shorrocks, 1997). If one assumes typical soil densities of 1700 kg/m3 and a mixing depth of 20 cm (default values used in the EUSES model or European Union System for the Evaluation of Substances), an application rate of 1–2 kg B/ha results in an estimated soil concentration of 0.3–0.6 mg B/kg soil. Mortvedt et al. (1992) estimated soil concentrations of 0.16–2.0 mg B/kg soil for several crops with application rates of 0.45–5.7 kg/ha. The intentional application of borates to achieve such soil concentrations should be acknowledged in any risk assessment process.

94.4.2  Biological Importance in Animals (Vertebrates) Boron has been found to be critical for normal reproduction and embryonic development in several animal species,

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and mechanisms for this essentiality are beginning to be revealed (Fort, 2002). Like many essential elements, it is likely that boric acid exhibits a U-shaped dose–response curve in animals. A beneficial effect to fish at low concentrations was shown for carp and rohu (Raymond and Butterwick, 1992). Work with rainbow trout and zebra­ fish has shown that embryo and larval development was adversely affected in waters deficient in boron (Eckhert, 1998; Rowe and Eckhert, 1999; Rowe et al., 1998). Fort et al. (1998, 1999) reported that abnormal development in frog embryos (Xenopus laevis) was observed when larval stages were exposed to low boron conditions of 0.003 mg B/l or less. But boron does not appear to be essential for all species. Growth of vitamin D3-deficient chicks was stimulated by supplementation of boron (3 mg B/kg diet) in a low-B basal diet (Hunt and Nielsen, 1981). Boron supplementation in pig diets (5 mg B/kg diet) decreased the inflammatory response to an intradermal injection of phyto­ hemagglutinin in pigs, altered plasma lipid metabolites, and tended to increase the production of cytokines following a stress (Armstrong et al., 2000, 2001; Armstrong and Spears, 2003). In rats, maternal exposure to a low-boron diet was associated with a reduction in embryo implantation sites (Lanoue et al., 1998b). In vitro exposures of mouse embryos to low-B growth medium showed reduced blastocyst formation and increased embryo degeneration (Lanoue et al., 1998a).

94.4.3  Biological Importance in Humans An extensive database exists relating to the nutritional importance of boron in humans. Several authors have proposed a role for boron in the metabolism of vitamin D and estrogen (Nielsen, 1998; Nielsen and Penland, 1999; Samman et al., 1998). In addition, dietary boron deprivation studies in both rats and humans have consistently found an effect of boron intake on brain electrophysiology; and in humans on performance of task measuring eye–hand coordination, attention, and short-term memory (Penland, 1998). Although to date insufficient data are available to confirm essentiality in humans, the U.S. Food and Nutrition Board in 2001 published a Tolerable Upper Intake Level (UL) for boron of 20 mg/day, which confirms the nutritional importance for humans. Also, the UK Expert Group on Vitamins and Minerals (UK EVM, 2003) and the European Food Safety Authority (EFSA, 2004) also regarded boron as nutritionally important and determined an acceptable daily intake for boron (0.16 mg/kg/day). More recent evidence of the biological importance of boron comes from the discovery of a Quorum sensing cell–cell communication auto-inducer molecule containing a borate-sugar diester (Chen et al., 2002) and the B transporter membrane protein, BOR1, identified in

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plant roots of Arabidopsis thaliana (Takano et al., 2002, 2005, 2008). Further support in humans comes from the Kibblewhite et al. (1984) study, which reported a significantly higher incidence of esophageal cancer in a lowboron region, compared to an area with boron exposure. A case–control study also found no significant elevation in the risk of prostate cancer in an occupational cohort with boron exposure (Rooney et al., 1993; Zhang et al., 2001). Researchers using human prostate cells have reported a biochemical mechanism that may explain a potential role for boron in the inhibition of human prostate cancer cell proliferation (Barranco and Eckert, 2004; Barranco et al., 2007; Gallardo-Williams et al., 2003, 2004; Henderson et al., 2009).

94.5  Toxicokinetics 94.5.1  Absorption Boric acid and sodium borates given orally are readily and completely absorbed in humans and animals. Studies in animals include rats (Ku et al., 1991), rabbits (Draize and Kelly, 1959), sheep (Brown et al., 1989), and cattle (Owen, 1944; Weeth et al., 1981), as shown by the levels of boron in urine, blood, or tissues. In adult human volunteers given a single oral dose of 131 mg B (as boric acid dissolved in water), 94% of the administered dose was excreted in the urine over a 96-h period (Schou et al., 1984). Similar absorption was observed based on urinary excretion of boron in six volunteers drinking a curative spa water with a high boron content (daily dose of 102 mg B) for 2 weeks (Job, 1973). In another study, greater than 90% was absorbed in human volunteers taking in 3% boric acid in an aqueous solution or as a waterless emulsifying ointment spread onto biscuits (Jansen et al., 1984b). In a series of human volunteer studies conducted in the early 1900s in which large doses of boric acid were repeatedly administered, approximately 80% of an administered dose was recovered in the urine, while 1% was recovered in the feces (Wiley, 1904). Reports involving accidental human ingestion, particularly in infants, provide further evidence of oral absorption (Wong et al., 1964). Inhaled sodium borate dust is readily absorbed as demonstrated by the blood and urine levels among groups of workers occupationally exposed to various levels of boron (Culver et al., 1993, 1994a). In rats, inhaled boron oxide aerosol was readily absorbed, based on the increased levels of boron excreted in the urine following inhalation exposure (Wilding et al., 1959). Dermal absorption of borates across intact skin is insignificant in all species evaluated, including human newborn infants (Friis-Hansen et al., 1982), adult humans (Beyer et al., 1983; Hui et al., 1996; Wester et al., 1998), rabbits (Draize and Kelley, 1959), and rats (Nielsen,

Hayes’ Handbook of Pesticide Toxicology

1970). Borates have been demonstrated to penetrate damaged or abraded skin (Draize and Kelley, 1959; Nielsen, 1970; Stüttgen et al., 1982). However, the use of an ointment-based vehicle may prevent or reduce the absorption through diseased skin compared to an aqueous jelly-based vehicle (Nielsen, 1970; Stüttgen et al., 1982), although the results by Stüttgen et al. have a number of flaws and are therefore not conclusive. Skin absorption data were obtained in human volunteers that were dosed over a 900 cm2 area (30  30 cm) of their back with 10B-enriched boric acid or disodium tetraborate decahydrate (5% in aqueous solution) or disodium octaborate tetrahydrate (10% in aqueous solution). Twenty-four hours later, the residual dose was removed by washing. Absorption of boron was measured from the urine (Hui et al., 1996; Wester et al., 1998). For the purposes of risk assessment, a dermal absorption for borates of 0.5% was determined based on the mean percentage of dose absorbed plus the standard deviation. However, this is only applicable to small volumes landing on the skin and remaining in contact. For larger volumes on the skin, not all the substance will be available for absorption; therefore the rate of absorption (flux) is more appropriate.

94.5.2  Distribution There is no evidence of boron accumulation (Alexander et al., 1951; Culver et al., 1994a; Forbes et al., 1954; Forbes and Mitchell, 1957; Jansen et al., 1984a; Ku et al., 1991, 1993; Treinen and Chapin, 1991; Ward, 1987), although bone may contain higher levels than other tissues. Absorbed boron rapidly distributes throughout the body water in humans and animals. In a study of workers occupationally exposed to 10 mg/m3 of airborne di­sodium tetraborate decahydrate (0.22 mg B/kg/day), there was no progressive accumulation of boron in soft tissues during the working week as measured by blood and urine levels (Culver et al., 1993, 1994a). Similarly, Jansen et al. (1984a,b) concluded from pharmacokinetic studies of human volunteers that there was no tendency for boron to accumulate following a single i.v. dose of 600 mg of boric acid (approximately 105 mg B). Tissue levels of boron generally reached steady-state within 3–4 days among rats fed boric acid in the diet or drinking water for 28 days (Treinen and Chapin, 1991) or 9 weeks (Ku et al., 1993). Thus, boron does not accumulate in soft tissues with time in either humans or animals. In both humans and animals, boron levels in soft tissue are comparable to plasma levels, while a greater concentration of boron in bone is observed relative to other tissues. The most complete study of boron distribution conducted to date examined tissue disposition of boron in reproductive organs and other selected tissues in adult male rats fed boric acid at approximately 100 mg B/kg bw/day for

Chapter | 94  Boric Acid and Inorganic Borate Pesticides

up to 7 days (Ku et al., 1991, 1993). All tissues examined, except bone and adipose tissue, appeared to reach steadystate boron levels by 3–4 days. Bone achieved the highest concentration of boron (two to three times plasma levels), and bone boron levels continued to increase throughout 7 days of dietary administration (Ku et al., 1991). In contrast, adipose tissue concentration was approximately 20% of the plasma level. No other tissues showed any appreciable accumulation of boron over plasma levels. In dogs, an accumulation in the brain, liver, and fat was reported after a high single dose of 2000 mg/kg bw boric acid (Pfeiffer et al., 1945). But the accuracy of the analytical procedures raises questions about this finding. Other studies have also shown a greater concentration of boron in bone relative to other tissues in humans (Alexander et al., 1951; Forbes et al., 1954) and rats (Forbes and Mitchell, 1957). Boron levels in a number of tissues have been measured (Abou-Shakra et al., 1989; Ciba and Chrusciel, 1992; Minoia et al., 1990, 1994; Sabbioni et al., 1990; Shuler et al., 1990; Ward et al., 1987). In mice, boron distribution appeared to be homo­geneous in the tissues examined, except for higher levels in the kidney (bone was not analyzed) (Locksley and Sweet, 1954; Laurent-Pettersson et al., 1992), but higher levels were found in bone in another study (Massie et al., 1990). In vivo and in vitro studies indicate that boric acid has a strong affinity for cis-hydroxyl groups. This may explain the higher concentrations of boric acid in bone, due to the binding to the cis-hydroxyl groups of hydroxyapatite.

94.5.3  Metabolism Boric acid is not metabolized in either animals or humans, owing to the high energy level required (523 kJ/mol) to break the B–O bond (Emsley, 1989). Other inorganic borates convert to boric acid at physiological pH in the aqueous layer overlying the mucosal surfaces prior to absorption. Additional support for this derives from studies in which more than 90% of administered doses of inorganic borates are excreted in the urine as boric acid. Boric acid is a very weak and exclusively monobasic acid that is believed to act, not as a proton donor, but as a Lewis acid, i.e., it accepts OH−. Because of the high pKa, regardless of the form of inorganic borate ingested (e.g., boric acid, disodium tetraborate decahydrate or boron associated with animal or plant tissues), uptake is almost exclusively (98%) as undissociated boric acid.

94.5.4  Excretion In both humans and animals, boron is excreted in the urine regardless of the route of administration. It is excreted with a half-life of less than 24 h in humans and animals.

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Boric acid is slowly eliminated from bone. In humans, 99% of a single i.v. dose of boric acid was excreted in the urine; the plasma half-life was calculated to be 21 h using a three-compartment toxicokinetic model (Jansen et al., 1984a). Following oral intake of an aqueous solution of boric acid, the urinary recovery was 94% (Jansen et al., 1984b); more than 50% of the oral dose was eliminated in the first 24 h, consistent with the 21-h half-life in the i.v. study. In a boron balance study, Sutherland et al. (1998) showed that only 8% of dietary boron is excreted in feces. Half-lives of 28.7 h and a mean of 13 h have also been reported from various poisoning cases (Astier et al., 1988; Litovitz et al., 1988). Elimination half-lives for animals have not been stated explicitly in the scientific literature, but they can be calculated or estimated from the data in the literature. In mice, assuming first-order kinetics for elimination, the half-life was estimated to be approximately 1 h, and in rat less than 12 h (Farr and Konikowski, 1963; Ku et al., 1991, 1993). In rabbits, 50–66% of an orally administered dose of boric acid was excreted in the urine in the first 24 h after dosing (Draize and Kelley, 1959). A recent study indicated that the half-life may be only 3 h (Vaziri et al., 2001) in both pregnant and nonpregnant rats. The major determinant of boric acid excretion is expected to be renal clearance since boric acid is excreted unchanged in the urine. Rats and mice generally have faster rates of renal clearance than humans since the glomerular filtration rates as a function of body mass are generally higher in rats and mice than in humans. Clearances of 40.4  3.2 ml/min/ 1.73m2 for disodium tetraborate decahydrate in male rats and 40 ml/min/1.73m2 for boron in mice (Farr and Konikowski, 1963; Usuda et al., 1998) have been reported, although there are methodological and/or analytical limitations in both studies. In more recent studies, boric acid clearance rates in nonpregnant rats and pregnant rats ranged from 29.0  5.7 to 31.0  4.5 and from 32.2  5.1 to 35.6  5.7 ml/min/ 1.73m2, respectively (Vaziri et al., 2001). In humans, Jansen et al. (1984a) determined a clearance rate of 55 ml/min/1.73m2 following an i.v. dose of 600 mg of boric acid (105 mg B). Farr and Konikowski (1963) also reported a similar value of 39 ml/min/1.73m2 in humans given 35 mg B/kg intravenously as sodium pentaborate, although there are methodological and analytical limitations to this 40-year-old study. In a more recent study, renal clearance rates in humans were 68.30  35.0 ml/min/ 1.73m2 for pregnant subjects and 54.31  19.35 ml/min/ 1.73m2 for nonpregnant subjects (Pahl et al., 2001). This indicates about 20–25% greater clearance in pregnant humans. A comparison of the renal clearance between rats and humans in terms of body surface area suggests that humans clear boric acid slightly faster than rats (1.7–1.9 times as fast), while a comparison by body weight indicates that humans may clear boric acid more slowly than

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rats (approximately three to four times slower). This apparent divergence in comparative clearances based on body weight vs. surface area between the two species is related to the high surface-to-mass ratio in the rats as compared to the humans and would suggest that there is little ­ difference in the clearance rates (Pahl et al., 2001; Vaziri et al., 2001).

94.6  Toxicology 94.6.1  Efficacy (Invertebrates and Fungi) Boric acid and sodium borate salt compounds prevent pest infestation by inhibiting pest metabolism on a ­ cellular basis. For instance, the tetrahydroxyborate anion forms a chelate complex with the cis-adjacent hydroxyl groups in the ribose sugar of nicotinamide adenine dinucleotide (NAD), with the cationic nitrogen of the nicotinamide moiety providing electrostatic stabilization of the chelate. In this configuration, NAD and NADP cannot be used by the dehydrogenase enzymes of glycolysis, the pentose phosphate pathway or the tricarboxylic acid pathway, and so the cellular energy-generating mechanisms, including adenosine triphosphate production, are shut down (Lloyd, 1998).

94.6.2  Laboratory Studies 94.6.2.1  Acute Studies Although available acute oral studies are not of modern standards and were performed prior to the introduction of GLP practices, they are reproducible across a number of studies and species, and of acceptable quality. In aqueous solutions at physiological and acidic pH, low concentrations of simple borates such as boric acid, disodium tetraborate decahydrate, and disodium octaborate tetrahydrate will predominantly exist as undissociated boric acid. The toxicokinetic and toxicological effects of boric acid, disodium tetraborate decahydrate, and disodium octaborate tetrahydrate are likely to be similar on a boron equivalents basis. Therefore, the data obtained from studies with the borates can be read across in the human health assessment for each individual substance. (a)  Acute Oral Toxicity Boric acid and inorganic borate salts are generally considered to be of low acute oral toxicity in mammals, including rats, mice, and dogs. No substantial differences in acute toxicity were seen in rats, mice, and dogs in the limited studies available. However, dogs exhibit an emetic effect in response to high doses of borates. The LD50 in dogs was determined to be 3980 mg boric acid/kg and 6150 mg disodium tetraborate decahydrate/kg (administered in a

Hayes’ Handbook of Pesticide Toxicology

capsule). The dogs vomited shortly after treatment at all doses (158 mg boric acid/kg and 246 mg disodium tetraborate decahydrate/kg were the lowest doses tested). No other adverse symptoms were seen (Keller, 1962; Weir and Fisher, 1972). The main symptoms of toxicity seen in all species tested were CNS depression, ataxia, and convulsions. Humans, which are reviewed below, display different acute symptoms compared with most animals. The acute oral LD50 of zinc borate is greater than 10,000 mg/kg (limit of tested dosages) in albino rats (Daniels, 1969). (b)  Acute Inhalation Toxicity Low acute inhalation toxicity was observed for those borates tested; the 4-h LC50 being 2 mg/l for boric acid, disodium tetraborate decahydrate, disodium tetraborate pentahydrate, and disodium octaborate tetrahydrate (Wnorowski 1994a,b,c,d, 1997). The samples were ground in a ball mill for 24 h to ensure that the substance was in a respirable form and the top dose tested, approximately 2 mg/l, was the highest level that was obtainable under the test conditions. Animal observations were limited due to the accumulation of test material on the walls of the exposure chamber. No acute inhalation toxicity data are available for the form of zinc borate used in pest control products, but acute inhalation LD50 (4-h) results are available from a noseonly test on a similar zinc borate compound, found to be 4.95 mg/l (Blagden, 1996). (c)  Acute Dermal Toxicity As anticipated because of their minimal skin ­ absorption, the acute dermal toxicity was low for those borates tested, i.e., LD50s were 2000 mg/kg for boric acid, boric oxide, disodium octaborate, disodium tetraborate decahydrate, and disodium tetraborate pentahydrate, and 10,000 mg/kg (limit of tested dosages) for zinc borate (Daniels, 1969; Doyle, 1989; Reagan and Becci, 1985; Vernot et al., 1977; Weiner et al., 1982).

94.6.2.2  Chronic Studies (a)  Genotoxicity A number of in vitro mutagenicity studies have been conducted with boric acid, disodium tetraborate decahydrate, or disodium octaborate tetrahydrate, including bacterial mutation assays in Salmonella typhimurium and Escherichia coli, gene mutation in mammalian cells (L5178Y mouse lymphoma, V79 Chinese hamster cells, C3H/10T1/2 cells), bacterial DNA-damage assay, unscheduled DNA synthesis (hepatocytes), chromosomal aberration and sister chromatid exchange in mammalian cell (Chinese hamster ovary, CHO cells). No evidence of mutagenic activity was observed (Bakke, 1991; Haworth et al., 1983;

Chapter | 94  Boric Acid and Inorganic Borate Pesticides

Landolph, 1985; NTP, 1987; Stewart, 1991). In addition, no mutagenic activity was seen in vivo in a mouse bone marrow micronucleus study on boric acid (O’Loughlin, 1991). (b)  Carcinogenicity In long-term feeding studies with boric acid and disodium tetraborate decahydrate in both rats and dogs, no carcinogenic effects were observed (Weir and Fisher, 1972). Other observations for the rats included lower food consumption, retarded body weight gain, coarse hair coats, haunched position, swollen pads, inflamed bleeding eyes, and changes in hematological parameters at the highest doses (58.5 mg B/kg bw/day). The National Toxicology Program (NTP, 1987) also evaluated boric acid but no carcinogenic effects were observed at doses of 75 mg B/kg bw/day and 200 mg B/kg bw/day. Effects on survival rate and reduced body weight gain were at the high doses. Testicular effects were noted in both studies (NTP, 1987; Weir and Fisher, 1972) which are discussed in more detail below. (c)  Reproductive Toxicity Effects on the testis have been observed in both subchronic and chronic studies in three species: rats, mice, and dogs. In rats, a single dose of 175 mg B/kg bw was found to cause reversible disruption of tubular spermiation (Linder et al., 1990), although no such effects were observed after a single dose of 350 mg B/kg (2000 mg boric acid/kg) (Bouissou and Castagnol, 1965). Effects tend to be similar in all three species, although most data derive from rat studies. The reproductive effects in rats at lower doses and shorter time periods start with reversible inhibition of spermiation. Early effects can be seen after 14 days of treatment, at doses around 39 mg B/kg (217 mg boric acid/kg bw/day), but at a lower dose of 26 mg B/kg (149 mg boric acid/kg bw/day) the effects take about 28 days to manifest (Ku et al., 1993). Higher doses lead to testicular atrophy, degeneration of seminiferous tubules, reduced sperm count, and a reduction in fertility, as seen in a three-generation study of boric acid and disodium tetraborate decahydrate in rats at 58.5 mg B/kg bw/day (NOAEL, 17.5 mg B/kg bw/day) (Weir, 1966a,b,c; Weir and Fisher, 1972). Similar results were seen in a 2-year study of boric acid and disodium tetraborate decahydrate at 58.5 mg B/kg bw/day (NOAEL, 17.5 mg B/kg bw/day) (Weir, 1966c,d,e; Weir and Fisher, 1972). In male rats fed disodium tetraborate decahydrate for either 30 or 60 days at 100 or 200 mg B/kg bw/day (NOAEL 50 mg B/kg bw/day), testis weight was reduced, testicular germ cells were depleted, selected testicular enzymes were affected, and fertility was reduced (Lee et al., 1978). As might be expected, while recovery from inhibition of spermiation occurred at the lower doses, there was no recovery from testicular atrophy when germ cells were lost.

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Fewer data are available for mice and dogs, but the results agree with the findings in rats. In a continuous breeding study of boric acid in mice, a dose-related effect on the testis (testicular atrophy and effects on sperm motility, morphology, and concentration) was noted; fertility was partially reduced at 111 mg B/kg bw/day and totally reduced at 221 mg B/kg bw/day. Effects on females were minimal. The NOAEL was 27 mg B/kg bw/day (154 mg boric acid/kg bw/day), although at this dose the motility of epididymal sperm was slightly affected without any effect on fertility (Fail et al., 1991). These findings are consistent with those in rats. Data in dogs are derived from two very limited 2-year feeding studies. Unfortunately, the published study does not reflect the information detailed in the original study reports (Weir, 1966d,e,f, 1967a,b; Weir and Fisher, 1972). In addition, Weir and Fisher (1972) estimated the dietary intakes from standard intake figures. However, the original study reports contained the actual dietary intake allowing for a more accurate measurement of the boron intake. These amended figures are used here. Groups of dogs were fed either boric acid or disodium tetraborate decahydrate at doses up to 10.2 mg B/kg bw/day (62.4 mg boric acid/ kg bw/day and 84.7 mg disodium tetraborate decahydrate/kg bw/day) in one study and 39.5 mg B/kg bw/day (233.1 mg boric acid/kg bw/day and 373.2 mg disodium tetraborate decahydrate/kg bw/day) in a second study. Only four male dogs per group were used in each study, and animals were sacrificed at various time periods such that observations were reported on groups of only one or two animals. One boric acid-treated and one disodium tetraborate decahydrate-treated dog were allowed to recover for 3 weeks. Some recovery was observed in each dog. Minor histopathological changes such as decreased spermatogenesis remained, which was less obvious in the disodium tetra­ borate decahydrate-treated dog. At 39.5 mg B/kg bw/day, testicular atrophy was observed; however, the effects in the only one disodium tetraborate decahydrate-treated dog investigated at 38 weeks were less severe than those seen in the control dog. Also, testicular atrophy was present in three out of four control dogs, so that the significance of the effect in the treated animals is difficult to assess. The NOAEL was determined to be 10.2 mg B/kg bw/day. While these data should be considered inadequate for risk assessment, they do confirm the effects and NOAEL seen in other species. (d)  Developmental Toxicity Only boric acid has been tested in developmental studies. Effects were observed at high doses in rats, mice, and rabbits. The majority of studies have been carried out in rats. In two separate studies conducted in the same laboratory, rats received a large number of dose levels (approximately 3.3, 6.3, 9.6, 13.7, 25, 28, and 59 mg B/kg bw/day on

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g­ estation days 0–20 and 94 mg B/kg bw/day on gestation days 6–15) in feed. The NOAELs for maternal toxicity and developmental effects were 13.7 mg/kg bw/day and 9.6 mg B/kg bw/day, respectively. A reduction in food intake and an increase in relative liver and kidney weight and a reduction in maternal body weight gain at higher doses indicated maternal toxicity. At nonmaternally toxic doses, there was a reduction in fetal weight and some skeletal anomalies, which, with the exception of shortened 13th rib, had reversed by postnatal day 21 at 13.7 and 28.6 mg B/kg bw/day in a study designed to look at postnatal recovery (Price et al., 1990, 1996). At higher maternally toxic doses, other indications of developmental effects were observed, including resorptions and visceral malformations (enlarged lateral ventricles; cardiovascular effects; anophthalmia and microphthalmia, and short and curly tails). However, these are likely to have been secondary to the maternal toxicity (Heindel et al., 1992; Price et al., 1990, 1996). Similar findings were observed in mice receiving estimated doses of 0, 43, 79, and 175 mg B/kg bw/day on gestation days 0–20 in feed. Maternal toxicity was indicated by mild renal lesions and at the highest dose increases in the relative kidney weight and food intake. A NOAEL was not determined for maternal toxicity. The key developmental effects observed were similar to those seen in rats, i.e., a reduction in fetal body weight at the mid dose (79 mg B/kg) and an increase in skeletal variations and malformations (missing lumbar vertebrae, fused vertebral arches, and short rib XIII) and resorptions at the highest, more maternally toxic dose. The NOAEL for developmental effects in mice was 43 mg B/kg bw/day (Heindel et al., 1992), but this dose was also a maternally toxic dose. In rabbits receiving estimated doses of 0, 11, 22, and 44 mg B/kg bw/day by gavage on gestation days 6–19, maternal toxicity was indicated by effects such as an increase in relative kidney weight, increased food intake, vaginal bleeding, and an increase in corrected weight gain. Developmental effects were seen only at the top dose, where the majority of the embryos were resorbed and malformations were primarily visceral (major heart and/or great vessel defects); however, these effects are likely to be secondary to the maternal toxicity. The only skeletal effect observed was a decreased incidence of rudimentary extra rib on L1, which was not considered biologically significant. The NOAEL for both maternal and developmental toxicity in the rabbit was 21.8 mg B/kg bw/day (Price et al., 1991).

94.7  Accidental poisionings 94.7.1  Animals One should not become complacent in handling borates just because these materials are considered to be of low acute toxicity. For example, a herd of 85 mixed-breed

Hayes’ Handbook of Pesticide Toxicology

beef cows weighing 350–400 kg were allowed to forage in a recently harvested peanut field along the edge of which a bag of borate fertilizer had inadvertently been left. The fertilizer bag was completely torn apart and 26 cows subsequently died (Sisk et al., 1988). Clinical signs were weakness, mild ataxia, general depression, and mild shivering in the neck, shoulder, and hindquarter muscles. Two cows suffered seizures. Greenish diarrhea and dehydration were observed in most affected cattle before death. A group of Hereford heifers was investigated for tolerance to elevated boron levels in drinking water (Green and Weeth, 1977). When given a choice, the heifers discriminated against concentrations greater than 29 ppm boron (added as disodium tetraborate decahydrate) and rejected concentrations above 95 ppm. None chose boron water over tap water. The safe tolerance was not determined in this study, but was proposed to be between 40 and 150 ppm boron. In a study in yearling heifers, the boron status of cattle was found to correlate with plasma and urine concentrations (Weeth et al., 1981). Another study was designed to produce toxicosis in goats and to study blood and cerebrospinal fluid parameters at selected dose periods (Sisk et al., 1990). Numerous blood factors were affected. There was evidence of CNS activity, including seizures.

94.7.2  Humans There is a large database of accidental or intentional poisoning incidents for humans. Many were the result of accidental use as an antiseptic for irrigating body cavities, treating wounds, or as a treatment for conditions such as epilepsy. Such medical uses are now obsolete. Historical accidental misuses also include the preparation of baby formula (1–14 g boric acid in the formula) and the topical use of pure boric acid powder for infants, which led to poisonings. This database is reviewed in several reports from poisoning centers as well as a detailed review of cases in the literature from the mid 1800s to the 1970s ( Goldbloom and Goldbloom, 1953; Kliegel, 1980; Litovitz et al., 1988; Valdes-Dapena and Arey, 1962; Wong et al., 1964). Humans display different acute symptoms compared with most animals. In the literature, the human oral lethal dose is regularly quoted as 2–3 g boric acid for infants, 5–6 g boric acid for children and 15–30 g boric acid for adults. These data are largely unsubstantiated. In most cases, it is difficult to make a good quantitative judgment, particularly since medical intervention occurred in most cases and there were often other unrelated medical conditions (Culver and Hubbard, 1996). Of 784 more recent reports of accidental ingestion, none were reported as fatal and 88.3% were asymptomatic. The estimated dose range was 10 mg to 88.8 g (Litovitz et al., 1988). However, a single intake of 30 g of boric acid was fatal in one case (Yoshitaka et al., 1993). Symptoms of acute effects may include nausea, vomiting, gastric discomfort, skin

Chapter | 94  Boric Acid and Inorganic Borate Pesticides

flushing, excitation, convulsions, depression, and vascular collapse. Humans with multiple exposures (high levels 1 g) show various symptoms that may appear singly or together, and include dermatitis, alopecia, loss of appetite, nausea, vomiting, diarrhea, and focal or generalized central nervous system irritation or convulsions. Many data come from the mid 1800s to around 1940, when boric acid and disodium tetraborate decahydrate were used systematically for a variety of medical conditions including amenorrhea, malaria, epilepsy, urinary tract infection, and exudative pleuritis (Kliegel, 1980). Daily oral doses in adults ranged from 1–14 g per day and repeated doses of 6–10 g/day were given for as long as several weeks. Doses greater than 3–5 g/day regularly caused vomiting and/or diarrhea in the first instance often accompanied by dermatitis and appetite suppression. As the dose became higher and the dosing period longer, symptoms included alopecia, disseminated maculopapular eruption followed by widespread desquamation, focal or generalized central nervous system irritation, and convulsions. The dermatitis, nausea, diarrhea, and vomiting symptoms also occurred in some patients receiving doses of 2 g boric acid/day (29 mg boric acid/kg bw/day) and above. In one such case, reduction of the dose from 2 g/day of boric (29 mg boric acid/kg bw/day) acid to 1 g/day (14 mg boric acid/kg bw/day) resulted in resolution of the effects (vomiting and dermatitis). In all cases where withdrawal of treatment was reported, recovery occurred with no lasting effects. The lowest recorded adult dose causing symptoms was 2 g/day boric acid (Kliegel, 1980). In children, where low levels can be estimated (Gordon et al., 1973; O’Sullivan and Taylor, 1983), infants aged from 6 to 19.5 weeks ingested borax (as a honey-borax mixture that had been applied to pacifiers) for periods up to 12 weeks. The intake in one case was 125 g of disodium tetraborate decahydrate (borax) over a 12-week period, while the other case reported by Gordon et al. was 9 g of boric acid over a 5-week period. Observed effects, which disappeared upon withdrawal of exposure, related to the CNS such as convulsions, generalized seizures, and focal seizures. There were no dermal effects. Minor occurrences of vomiting and loose stools were also described.

94.8  Findings from studies   with humans 94.8.1  Occupational Studies Estimates of occupational exposures to borates usually refer to the large number of workers believed to be occupationally exposed. The National Occupational Exposure Survey (NOES) collected survey information in 1981–1983 based on site visits to 4490 establishments, identifying nearly 13,000 agents in over 100,000

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trade-name products (CDC/NIOSH 2009). Numerous borate products were noted with large numbers of potentially exposed workers. Estimated worker exposure to boric acid (CAS#10043-35-3, Agent Code X7080) was reported for 489,668 workers. Curiously, the number exposed to “boric acid” (no CAS number stated, Agent Code 12960) was 301,748. Observations of 117 pest control workers resulted in reported exposure to boric acid. Exposures to registered nurses represented about 25% of the occupationally exposed individuals, machinery workers about 7.5%, and janitors and cleaners about 5% of the occupational exposures. Such values were used to prioritize borates for further studies. Moorman et al. (2000) cited 492,618 workers exposed to boric acid in their prioritization report. Garabrant et al. (1984, 1985) published the results of a study sponsored by NIOSH among the mine and refining workers employed at a Boron, California facility. Participants were asked standardized questions on respiratory disease and additional (nonstandardized unspecified) questions about other symptoms. Pulmonary function tests and chest x-rays were also administered. Symptoms affected included dryness of the mouth, nose, or throat, dry cough, nose bleeds, sore throat, productive cough, shortness of breath, and chest tightness. This included a reported significant increase in symptoms between 4 and 14.6 mg/m3 expressed as total dust. Based on these results, Garabrant et al. recommended that occupational exposures be kept below about 4 mg/m3 total particulates. However, exposure information was from historical records rather than measurements concurrent with symptom questioning and thus uncertainty exists about the reported relationship between exposures to borate dust greater than 4 mg/m3 and symptoms of exposure. In a later evaluation of the same worker population, Wegman et al. (1994) concluded that changes in pulmonary function over the 7 years since the Garabrant et al. (1984, 1985) studies were unrelated to any estimate of cumulative dust exposure. To evaluate acute irritation, Wegman et al. asked participants to report irritation events while a technician accompanied the participant and monitored exposure. Symptom questionnaires were administered before work and at hourly intervals throughout 4 consecutive workdays. Severity of symptoms was also noted on a scale of 0 to 10, beginning at grade 0 (“not at all”) and ending above grade 10 (“maximal”). Symptoms included nasal irritation, eye irritation, throat irritation, cough, and breathlessness. Exposures among the comparison group averaged 0.45 mg/m3 total dust with almost 100% of the samples below 1.0 mg/m3 total dust. Average daily exposure for the exposed workers was 5.72 mg/m3 total dust with 68% of the subject-days including one or more exposure above 10 mg/m3 total dust (the 15-min occupational limit). Based on the significance of exposure–response trends (probability of a symptom with increasing dust exposure), Wegman et al. concluded that irritation of the eye, nose,

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and throat, cough, and breathlessness did increase with dust ­ exposure. Wegman et al. concluded that the totality of evidence suggested that current boron exposures at the facility monitored were generally associated with no more than moderate irritant responses and do not indicate chronic, irreversible changes. They acknowledge that the results could be used to set an acceptable level of risk for a sensory irritant. Using the probability of any symptom occurring, they concluded that a daily level of 10 mg/m3 total dust would make more than one moderate symptom per day improbable. Cain et al. (2004) evaluated sensory perceptions of borate dusts, in comparison with other mineral dusts (calcium oxide, calcium sulfate). Subjects were asked to rate mineral dusts based on their own sensitivity to a series of carbon dioxide concentrations; using a scale corresponding to CO2 concentrations provided a way to standardize the individual perceptions of sensory irritation. Twelve subjects were exposed to dust particles for 20 min while performing moderate exercise (i.e., riding an exercise bike set at a load of 60 watts). During exposure, subjects judged the level of feel or irritation in the eye, nose, and throat (nasopharynx) at 5-min intervals. Nasal resistance, nasal secretion, mucociliary transport, heart rate, oxygen saturation, minute ventilation, and respiration rate were recorded as well. Subjects did not know the identity or concentration of the test substances, which were presented in 14 different sessions. The comparative responses to the substances indicated that CaO was about five times more potent than disodium tetraborate pentahydrate. In no instance did the subject wish to stop the exposure because of the severity of the perceived irritation, suggesting that no exposure evoked a severe response. Subjects could perceive all the dusts, but at different concentrations. Subjects perceived CaO above background at 1, 2, and 5 mg/m3 in the nose and throat, but not in the eye. Subjects perceived sodium borate at 5 mg/m3 in the nose, but only at 10 mg/m3 and higher for the nose and eye. Nasal secretion increased above background at 10 mg/m3 for sodium borate. No dust-related effects were seen in mucociliary transport, nasal resistance, or other physiological measurements. The relative responses to the dusts suggest that the OEL for sodium borates should be about five times that of CaO, which is consistent with the 10 mg/m3 value used with nuisance dusts not otherwise classified. A second study (Cain et al., 2008) included boric acid among the mineral dusts and extended the duration of exposure. Responses to boric acid were slightly less than to sodium borate, although not significantly so. The magnitude of response confirmed that peak response occurred at about 20–30 min, either decreasing or plateauing through the 45-min exposure. Cain et al. reported that the highest dust levels studied were at the edge of where people considered carbon dioxide to become irritating (17–18% CO2)

Hayes’ Handbook of Pesticide Toxicology

and so the concentrations studied would be unlikely to cause irritation in an occupational setting. Boric acid is also added to some metalworking fluids as a biocidal agent. Krystofiak and Schaper (1996) used an animal test to evaluate the potential contribution of boric acid to the sensory and pulmonary irritation of a semisynthetic metalworking fluid. Although the metalworking fluid and several components were found to evoke irritation responses, the boric acid (and boramide, a reaction product of boric acid with amines in the metalworking fluid) did not evoke significant responses and would therefore be unlikely to contribute to sensory irritation in the workplace. In sum, occupational studies to borate dusts suggest that these substances can be perceived in the workplace as sensory irritants, primarily of the nose. However, the severity of response has been found to be low. Maximal responses to dust are likely to occur within tens of minutes of exposure and did not result in any residual effects over years of workplace exposure.

94.8.2  Epidemiological Studies Three major studies of the reproductive health among boron-exposed groups have been reported. These were published by Whorton and colleagues based on studies at a California (USA) mine (Boron, California, USA); by Sayli based on studies of families living in boron-rich areas of Turkey, and by Robbins, Wei and colleagues based on studies at Chinese sites, comparing workers in the boron industry with others living in the local community (naturally high in boron) and a distant community. An additional study has been reported among Russian workers, but the reliability of this study should be regarded as poor (Tarasenko et al., 1972). Furthermore, workplace dust exposures were reported to be four to eight times the 10 mg/m3 Russian permissible exposure limit. Whorton et al. (1994a,b, 1992) investigated reproductive health effects among 753 workers at the U.S. Borax (now Rio Tinto Minerals) mine (Boron, California). Whorton et al. reported no significant adverse effects from exposures to borates, specifically evaluating fertility compared with U.S. national averages. Five exposure categories were constructed to contain equivalent numbers of subjects. Dust exposures were based on previous studies of exposures associated with job categories [Exposure categories were based on Hu et al. (1992), who obtained samples from 1981–1988 and characterized exposures by job titles.], as shown in Table 94.3. Reproductive data were obtained by questionnaires and telephone interviews. To determine whether fertility was associated with occupational exposure to boron, Whorton et al. used the Standardized Birth Ratio based on a comparable nonexposed United States population. This incorporates ­adjustments for maternal age, race, parity, and time periods

Chapter | 94  Boric Acid and Inorganic Borate Pesticides

for the study population to the U.S. national data. Whorton et al. reported that there were 529 observed births fathered by the participants when only 466.6 births were expected. As shown in Table 94.3, there was no dose–response pattern to excess births and exposure category: the highest numbers of excess births were to the lowest and highest exposure categories. The incidences of reported infertility or couples seeking assistance for fertility problems were less than national averages. This excess number of births was statistically significant, indicating fertility among U.S. boron mine workers was not adversely affected. Whorton et al. evaluated boy/girl ratios as well. Overall, boron workers fathered 52.7% female offspring vs. the US national average of 48.8%. This observation is often given as evidence that boron exposure was associated with a reduction in male offspring. However, Whorton et al. noted that such a conclusion was inconsistent with the data. Workers in the two lowest exposure categories had the highest percentage female offspring, while workers in the highest exposure category had virtually the same percentage (49.2%) as the national average. Table 94.3 shows that the fraction of boys increases across mean dust exposures: the highest exposed categories had the largest ratio. A simple linear correlation of average boron dust exposure with percent of male offspring suggests a significant positive relationship (R2  0.72). Although the sample size is small and limits the generality of the conclusions, this pattern does not appear to demonstrate that boron exposure led to fewer boys, but rather, higher boron exposure correlated with more boys.

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Whorton et al. pointed out that there is an excess number of both male and female offspring. They offered no explanation why the number of excess female offspring is unexplainably large. They noted a relatively high rate of vasectomies, about five times the national average. This may have reduced the fertility rate. Whorton et al. noted that female offspring exceeded male offspring for 3 of the 4 decades studied: only in the 1960s were there more males than females. The study did not include a community control group of men not working at the mine site, which might have allowed a more appropriate baseline than comparing to the U.S. national averages. This limitation is frequently lost in summarizations of this study as demonstrating that boron exposure in mine workers reduced the ratio of boys. More recently, a review of the gender ratio in California reported a statistically significant decrease in the ratio of male:female offspring among white residents state-wide (Smith and Von Behren, 2005). This suggests that any pattern observed during the California boron mine worker study might be entirely consistent within the larger statewide pattern and not related at all to boron exposure. Whorton et al. also identified a subset of 42 subjects who had worked at high exposure jobs for more than 2 years. This group of workers was older than the general worker population and 48% had had a vasectomy. The observed number of births for this group was near the expected number (SBR  102) considering their entire work history, but was equivalent to other workers during the period when they were in high exposure jobs (SBR 115–121).

Table 94.3  U.S. Mine Worker Reproductive Health and Dust Exposuresa Exposure category

1

2

3

4

5

Number of respondents

108

108

108

109

109

Range of dust exposures (mg/m3)

0.82

0.82-1.77

1.77-2.97

2.97-5.04

5.04

Mean dust exposure (mg/m3)

0.37

1.34

2.23

3.98

8.58

Estimated daily boron exposure (mg-boron/day)b

0.45

1.64

2.73

4.88

10.5

Observed births

94

108

93

114

120

Number of excess births (observed – expected)

31.9

3.8

-1.1

3.8

24.1

% Male offspring

44.7%

43.7%

48.4%

49.1%

50.8%

% Vasectomy

29%

43%

34%

41%

35%

a

Data from Whorton et al., 1992, 1994a, 1994b Estimate based on mean dust exposure, assumed 14% boron concentration in sodium borates, respiratory volume of 8.75 m3 air/day and 100% absorption. Whorton et al. used these calculations to estimate occupational exposure for a subset of the study group but did not estimate daily boron exposure for these categories.

b

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Thus, fertility was not adversely affected when they were most exposed to boron. The ratio of boys to girls was not reduced either and was equivalent to the ratio for the rest of the participants. A calculated daily boron exposure is shown for the other exposure categories in Table 94.3. Assuming dietary intake of boron is in the range of 1–3 mg boron/day, then total daily boron for most workers at the mine would not exceed the WHO (1998) daily upper intake limit of 13 mg boron/ day. Culver et al. (1994a) reported that the blood-boron concentrations of the highest exposure group were still an order of magnitude below blood concentrations having no effects in animals. Sayli et al. (1998, 2001, 2003) and Tuccar et al. (1998) reported comparisons of reproduction from families living in regions of Turkey with varied boron concentrations in the environment. In addition, some of the subjects worked at borate mining and processing facilities in the high-boron region. They reported no significant differences among the 1068 families in the high-boron region and the 610 families in the low-boron region. Sayli et al. reported nonsignificant differences in gender ratio, with more females than males (52.73% female) in the boron-rich region than in the boron-poor region (48.86% female). Çöl et al. (2000) evaluated infertility rates, gender ratio at birth, numbers of stillbirths and spontaneous abortions, premature births, and infant mortality rates among the families of 799 boron-exposed workers at three production facilities in Turkey. Patterns were compared with national or regional values, and 642 production workers were compared with 157 office workers. No significant adverse effects were found. Infertility rates among the workers averaged 1.8%, at the low end of the Turkish national rate of 1.49–3.8%. The gender ratio was 1.12 (52.9% boys, 47.1% girls), which is higher than the Turkish national range of 1.05–1.08. When comparing the production workers (expected to have higher boron exposures) to office workers, the only significant differences were that average pregnancies and live births among production workers exceeded those of office workers. Çöl et al. therefore concluded that boron exposure to workers did not adversely affect any of the indicators of fertility or development studied, including gender ratio. However, the infertility rate appears significantly smaller than values usually reported in other countries and may make conclusions from this study difficult to interpret. The most recent series of studies reported in the literature involved groups of workers in Liaoning Province in northeast China. The intent of the research was to investigate potential adverse male reproductive health among boron workers using both traditional sperm assays as well as newer biochemical techniques. Workers at boron mining and processing plants located in Kuandian City comprised the high-exposure group. Residents who did not work in the boron industry but lived in the local area (which is

Hayes’ Handbook of Pesticide Toxicology

naturally high in soil boron) were included as a local community comparison group. Another population was selected from Tiantuia Gu, a town about 50 km away from the boron mines and processing plants with a low background of environmental boron and included as a distant control community group (Chang et al., 2006; Robbins et al., 2008). Liu et al. (2005) compared semen quality data for subjects with exposure to less than 13 mg boron/day (low exposure, n  148) with subjects exposed to more than 13 mg boron/day (high exposure, n  28). This criterion represents the WHO (1998) upper daily recommended limit for boron uptake. Boron exposure was determined based on end-of-shift urine boron concentration adjusted for creatinine concentration. There were no significant differences in sperm motile velocity, sperm motility, sperm concentration, total sperm, sperm production, or zinc concentrations. Liu et al. (2006a) compared semen quality between a boron exposure group, a community control (of non-boron workers) group, and a background control group. The stated conclusion was that, “under the current daily boron exposure dose, 13 to 430 mg/day, no negative reproductive effects on male workers are found as to semen quality.” Based upon 11 semen parameters, they used cluster analysis to develop four semen indices, representing sperm quantity, motility, speed of the motile cells, and departure of the sperm path from a straight line. No significant differences were found among the three exposure groups. Comparing the data with WHO standards, no statistically significant difference was reported. Gao et al. (2007) reported semen quality in 645 healthy men from six Chinese provinces, noting that 4.8% of the subjects had sperm concentrations less than the WHO standard, 7.1% had total sperm number less than the WHO standard, and 82% had a-grade sperm less than the standard. Liu et al. (2006b) compared 66 boron workers, 61 local community residents and 68 control area residents with regard to sperm density, motility, a-grade sperm, and total sperm and reported that there was no statistical difference among the groups. They also reported that the maximum exposure to boron was 430 mg/day, with 23 being exposed to more than 23 mg/day, 15 to more than 50 mg/day, and 10 higher than 100 mg/day. A significant negative impact of legume and dry fruit consumption on sperm quality was found. Liu et al. (2006c) reported data from this same study regarding sperm density, motility, a-grade sperm, and total sperm and reported that there was no statistical difference among the groups. They also compared data from subjects exposed to the WHO (1998) daily boron intake safe upper limit (13 mg/day) and reported no significant differences in these same parameters. Robbins et al. (2008) evaluated the ratio of sperm bearing Y-chromosomes to sperm bearing X-­chromosomes

Chapter | 94  Boric Acid and Inorganic Borate Pesticides

using the FISH technique (fluorescence in-situ ­hybridization) with samples from 146 men in the study and concluded that the ratio was significantly decreased among boron workers when compared to both men living in the area but not employed in the boron industry and to men living in an area of low environmental boron. They also concluded that the Y:X sperm ratio explained a decreased ratio of boy-to-girl offspring at birth among the highly exposed boron workers. Robbins et al. reported that the boron worker group had total daily boron exposures of 41.2  37.4 mg B/day. This was almost ten times the total exposure of the local nonboron workers (4.3  3.1 mg B/day) and about 18 times the total exposure of the control group (2.3  3.0 mg B/ day). The ratio of Y:X sperm was lowest among the highexposure boron worker group (0.93  0.03, 63 subjects) compared to the non-boron worker group (0.96  0.04, 39 subjects) and the control group (0.99  0.03, 44 subjects). Since gender is determined by the presence of a Y-chromosome (producing males) or an X-chromosome (producing females), a changed in the Y:X ratio could lead to an altered gender ratio at birth. A decrease in the Y:X ratio could suggest a mechanism to explain a decrease in the proportion of male children. To evaluate gender ratio, Robbins et al. did not report the standardized birth ratio (SBR) of boys to girls. Rather, they used a value termed “more boys than girls.” The study authors reported that 57.7% of the boron workers had fathered more boys than girls. The non-boron workers reported 42.3% fathers with more boys than girls, and the control group reported 76.7% fathers with more boys than girls. However, these analyses excluded all men with equal numbers of male and female offspring at birth. Robbins et al. did not indicate why the gender ratio estimate for the most highly exposed group (boron workers) was intermediate to the two lower exposed groups but emphasized the change in Y:X ratios. Whorton et al. (2008) evaluated the conclusions of Robbins et al., noting the lack of an exposure–response pattern between boron exposures and the endpoints, suggesting the potential for confounding factors. They also observed the relatively high incidence of elective abortions in both worker and community groups. Since the gender ratio in the community control group (1.76) is extremely high (even than the national Chinese ratio of 1.38 or the worldwide ratio of 1.02–1.05), it may present an unreliable baseline. No significant differences in “more boys than girls” were reported by Chang et al. (2006) from 938 boron mining and processing workers with a comparison group of 251. This was the same research group that authored the Robbins et al. (2008) publication, but Chang et al. concluded that there were no significant differences in frequency of more boys than girls (55.52% for boron workers vs. 60.31% for controls, p  0.234). Considering all offspring, 52.45% of the offspring to men in the boron group

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were males vs. 54.35% for the comparison group. No significant differences were also reported for multiple births, spontaneous miscarriages, stillbirths, tubal or ectopic pregnancy, or mean number of pregnancies fathered altogether. Chang et al. reported a trend (not statistically significant) in delay in pregnancy among boron workers, and slightly decreased average number of live births. Chang et al. (2006) did not report quantitative exposure measurements but calculated air dust exposure as extremely variable, ranging from 0.06 to 51.07 mg/day. They also reported environmental boron concentrations: surface water, 2.6–3.8 mg B/l; well water, 1.2–25.1 mg B/l; and soil, 133–1195 mg B/kg. They did observe that the use of respiratory protection was variable, with the halfface air-purifying respirators in some cases lacking filters and change of filters was not always frequent. They also reported that 64% of the boron workers ate at the worksite with as many as 33% eating directly in the work area. Xing et al. (2007) concluded that the risk of elevated boron exposure for workers from boron mining and processing areas was much higher than that of the other subjects. Drinking water concentrations in the mining area were 2.05 mg boron/l vs. 0.86 mg/l in the local community, and 0.05 mg/l in the background community control group. Food boron concentrations were 9.46 mg/kg (in the mining area) vs. 6.19 mg/kg and 3.29 mg/kg in the comparison communities. Daily intake from food and water was 8.00 mg/day (boron workers) vs. 4.25 mg/day and 1.40 mg/ day in the comparison communities. Workers drinking water from the staff canteen well and having dinner in this canteen had 95.5–469 mg/day boron exposure, with an average of 219 mg/day. Xing et al. (2008) reported correlations between total daily boron exposure and urine, as well as boron in blood, semen, creatinine-corrected post-shift urine and 24-h urine. They suggested that creatinine-corrected postshift urine could be used as a biomarker of worker exposure. They reported that the mean total daily boron exposure for the 60 boron workers evaluated in 2003 was 37 mg boron/day, with a geometric mean of 14.6 mg boron/day, both mean values clearly exceeding the WHO upper safe limit value of 13 mg boron/day. Among the 15 workers sampled in 2004, 87% of the boron workers with total exposure data had boron exposures exceeding the WHO (1998) upper safe limit, with a mean of 41.2 and geometric mean of 29.1 mg boron/day. The publications of the Chinese studies were reviewed by Scialli et al. (2009). They concluded that no clear evidence of male reproductive effects were attributable to boron, even including workers with a mean daily boron intake of over 31 mg B/day and a subset of workers with an estimated mean daily intake of 125 mg B/day. The review agreed that there were no statistically significant differences in semen endpoints and no differences in ­reproductive outcomes in the wives of the 945 boron workers and the wives

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of 249 background control men after adjustment for potential confounders, nor was the sex ratio found to be significantly altered in the boron exposure group. The review concluded the studies suggested a factor other than boron was responsible in the for the relationship between Y:X chromosome ratio and group status.

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Wong, L. C., Heimbach, M. D., Truscott, D. R., and Buncan, B. D. (1964). Boric acid poisoning: report of 11 cases. Canad. Med. Assoc. J. 90, 1018–1023. World Health Organization (WHO) (1998). “Guidelines for Drinking Water Quality”. 2nd ed. Addendum to Volume 1. Recommendations Boron, pages 4-6 adn Addendum to Volume 2 Boron pages 15-29. World Health Organisation, Geneva. Xing, X-u., Wu, G., and Wei, F. (2007). Daily intake of boron through foood and drinking water among people living in boron industrial area. J. Environ. Health 24(3), 125–127.

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Xing, X-u., Wu, G., Wei, F., Liu, P., Wei, H., Wang, C., Xu, J., Xun, L., Jia, J., Kennedy, N., Elashoff, D., and Robbins, W. (2008). Biomarkers of environmental and workplace boron exposure. J. Occup. Environ. Hygiene 5, 141–147. Yoshitaka, I., Fujizuka, N., Toshihiko, T., Shimizu, K., Tuchida, A., Yano, S., Naruse, T., and Chishiro, T. (1993). A fatal case of acute boric acid poisoning. Clin. Toxicol. 31, 345–352. Zhang, Z-F., Winton, J. I., Rainey, C., and Eckhert, C. D. (2001). Boron is associated with decreased risk of human prostate cancer. FASEB J. 15, A1089 (834.3).

Chapter 95

Imidacloprid: A Neonicotinoid Insecticide Larry P. Sheets Bayer CropScience, Research Triangle Park, North Carolina

95.1  Introduction Imidacloprid is a broad-spectrum neonicotinoid insecticide, with excellent systemic and contact activity that supports its use on many food crops, turf and ornamentals, and for termite and flea control. Imidacloprid is the most well known and widely used representative of the neonicotinoid insecticides. Neonicotinoid insecticides are designed to act on nicotinic receptors to control insect pests and, at the same time, to express low toxicity to vertebrate species. This is accomplished by selecting compounds for commercial development with high specificity for nicotinic receptor subtypes that occur in insects. The activity of neonicotinoid insecticides on the central nervous system of vertebrates is further reduced by poor penetration of the blood–brain barrier. The toxicology database supports the success of this strategy for imidacloprid, with signs of nicotinic stimulation (e.g., tremor) evident only at relatively high levels of exposure. By oral administration, imidacloprid is rapidly absorbed, metabolized in the liver, and excreted, primarily via the urine. Results from dietary-exposure studies support rapid metabolism, with little evidence of cumulative toxicity and minimal effects, even at high doses. Imidacloprid is not mutagenic in vivo or carcinogenic. Furthermore, it is not a primary embryotoxicant or a reproductive toxicant, nor is it teratogenic. Due to its high insecticidal potency and relatively low mammalian toxicity, imidacloprid has a very high margin of safety.

95.2  Historical overview 95.2.1  Chemistry Imidacloprid [1-[(6-chloro-3-pyridinyl)methyl]-N-nitro-2imidazolidinimine] is the first representative of the neonicotinoid insecticides that was registered for use and remains the most important commercial product. The history of the neonicotinoids began in the late 1970s, when Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

chemists at Shell Chemical Company investigated the heterocyclic nitromethylenes as potential insecticides (Schroeder and Flattum, 1984; Soloway et al., 1978). An excellent review of the discovery and early development of these insecticides has been compiled by Yamamoto and Casida (1999). The term “neonicotinoid” is used to distinguish these chemicals from the nicotinoids (Tomizawa and Yamamoto, 1993), with the neonicotinoids being more highly effective as insecticides and less toxic to vertebrate species. Representative insecticides are also referred to as “chloro­ nicotinyls” to emphasize the importance of the chlorine atom for insecticidal potency. Imidacloprid was discovered in 1984 by chemists at Nihon Bayer Agrochem who were exploring the introduction of a 3-pyridylmethyl group on the nitromethylene heterocycle structure (Shiokawa et al., 1986). The introduction of this moiety has been shown to greatly increase insecticidal activity and reduce mammalian toxicity (Kagabu et al., 1992; Zwart et al., 1992, 1994), while retaining the many other properties that are important for commercial applications, including photostability. Since the discovery of imidacloprid, several other chemical analogues with the 6-chloro-3-pyridylmethyl moiety have been developed for commercial use (Figure 95.1). Included in this group are acetamiprid (Takahashi et al., 1992; Yamada et al., 1999), nitenpyram (Minamida et al., 1993), and thiacloprid. The replacement of the chloropyridinyl moiety with a chlorothiazolyl group led to the development of a “second generation” of neonicotinoid insecticides (Maienfisch et al., 1999). This substitution has been shown to further reduce potency in assays with mammalian receptors but does not appear to reduce toxicity to mammals or to reduce activity at the insect nicotinic receptor (Chao and Casida, 1997; Liu et al., 1993; Zhang et al., 2000). Compounds in this group that have been developed for commercial use include clothianidin and thiamethoxam (CGA 293’343; Maienfisch et al., 1999) (Figure 95.1). Thiamethoxam was the first representative from this group to be registered for use (Wiesner and Kayser, 2000). 2055

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CHLORPYRIDINES Cl

Cl N

N

CH3 N

N

N

Acetamiprid

N

NH

NH

N

NO2

Imidacloprid Cl

S

NH

N

CHLOROTHIAZOLES

S CH3

N

N N

Thiamethoxam

CN

NO2

Clothianidin O

Cl N

CH3

CH3

NO2

CH3

Cl N

N Nitenpyram

NH

CH3

NO2

Cl N

N

S

N CN Thiacloprid Figure 95.1  Neonicotinoid insecticides.

Dinotefuran was the first representative of the furanicotinyl subclass of neonicotinoid insecticides. Representatives in this group have a ()-tetrahydro-3-furylmethyl moiety, using acetycholine as the lead compound instead of nicotine (Wakita et al., 2003). Although structurally distinct, representatives of the furanicotinyl subclass are considered neonicotinoid insecticides because they act through a similar mode of action, by binding to nicotinic acetylcholine receptors.

95.2.2  Nicotinic Activity 95.2.2.1  Insects The insecticidal activity of the neonicotinoids is attributed to actions on post-synaptic nicotinic receptors (Buckingham et al., 1997; Nagata et al., 1998), which, in insects, are located exclusively in the central nervous system. In insects, multiple subtypes of nicotinic receptor have been identified, which express different physiological and pharmacologic properties (Gundelfinger and Schulz, 2000; Wiesner and Kayser, 2000). With respect to the neonicotinoids, it has been determined that imidacloprid acts on at least three pharmacologically distinct subtypes of nicotinic receptor in the cockroach (Buckingham et al., 1997). Further characterization

of the nicotinic receptors that exist in insect tissues and the relative activity of neonicotinoids on the various subtypes continue to be active areas of research. The treatment of insect neuronal preparations with a neonicotinoid insecticide produces a bi-phasic response, consisting of an initial increase in the frequency of spontaneous discharge that is followed by a complete block to nerve impulse propagation (Schroeder and Flattum, 1984). Signs of intoxication in the American cockroach (Periplaneta americana) following exposure to imidacloprid consist of uncoordinated abdominal quivering, wing flexing, tremor, and violent whole-body shaking, followed by prostration and death (Schroeder and Flattum, 1984). Insecticidal activity is greatly enhanced by synergists that inhibit oxidative degradation (Liu and Casida, 1993), which would appear to support including a synergist in commercial formulations.

95.2.2.2  Mammals Mammalian tissues contain many subtypes of nicotinic receptor, which are derived from five homologous subunits, in combinations that are formed from nine , four , , , and  subunits (Tomizawa et al., 1999). In mammals, nicotinic receptors are located in many tissues, including autonomic ganglia, skeletal muscle (neuromuscular junction), spinal cord, and a number of brain regions. Differences in binding properties to the various receptor subtypes contribute greatly to the much lower activity of neonicotinoids in vertebrate tissues, as compared to tissues from insects (Yamamoto et al., 1998). An extensive database for differential sensitivity with imidacloprid (Chao and Casida, 1997; Liu and Casida, 1993; Matsuda et al., 1998; Methfessel, 1992; Nagata et al., 1999; Tomizawa et al., 1999) has been summarized by Tomizawa and Casida (1999). The relative specificity for the nicotinic receptor in insects is used to select compounds for commercial development. The success of this strategy is reflected by very high margins of safety for these insecticides (Leicht, 1993). The acute toxicity (defined by lethal potency) of various neonicotinoid insecticides and related analogues in mammals is most closely related to potency at the 7 nico­ tinic receptor subtype, with a decreasing relationship reported sequentially at 4, 2, 3, and 1 nicotinic receptors (Tomizawa and Casida, 1999). However, acute toxicity in mammals involves complex actions at multiple receptor subtypes, with relative subtype specificity conferred by minimal structural changes. Furthermore, the actions of neonico­ tinoids at these receptor subtypes involve a combination of agonist and antagonist activities (Nagata et al., 1998). Given this combination of actions, known or expected differences in relative specific activity at each nicotinic receptor subtype, and expected differences in distribution to target tissues, the toxic effects in vivo would likely vary among representative compounds. However, there has been no systematic assessment of toxicity in vivo, with tests conducted under appropriately standardized conditions.

Chapter | 95  Imidacloprid: A Neonicotinoid Insecticide

95.3  Metabolism and toxicokinetics The information available on the metabolism and toxico­ kinetics of imidacloprid in the rat is described in additional detail elsewhere (Thyssen and Machemer, 1999). Briefly, there are two major routes of metabolism in mammals. The first involves oxidative cleavage to imidazolidine and 6-chloronicotinic acid, with the imidazolidine moiety excreted via the urine. The nicotinic moiety is further degraded via glutathione-conjugation to a derivative of mercapturic acid and then to methyl mercaptonicotinic acid. This moiety is also conjugated with glycine to form a hippuric acid conjugate for excretion. The second substantive route in the biotransformation of imidacloprid involves the hydroxylation of the intact molecule in the imidazolidine ring, followed by the elimination of water and the formation of an unsaturated metabolite. In rats, there are no qualitative differences between males and females after the oral administration of a low dose of 1 mg/kg body weight or a dose of 20 mg/kg body weight. The same complement of metabolites is present in male and female rats at both dose levels, although at the higher dose of 20 mg/kg body weight, orally treated females exhibit a slightly higher renal elimination than males. More than 90% of a given dose is eliminated within 24 h, with total excretion by 48 h. Eighty percent of the dose is excreted via the urine, with the remainder of the dose eliminated via the feces. Imidacloprid is absorbed and widely distributed to organs within 1 h following oral administration to rats. Whole-body autoradiography indicates that imidacloprid is not distributed to fatty tissues, to tissues in the central nervous system (CNS), or to the mineral components of bone. These results indicate that there is low potential for accumulation and poor penetration of the blood–brain barrier, at least to dose levels of up to 20 mg/kg body weight. Poor penetration of the blood–brain barrier has also been reported with other neonicotinoids (Yamamoto et al., 1995). This property reduces their access to nicotinic receptors in the CNS, such that centrally mediated effects would not be expected at low levels of exposure.

95.4  Mammalian toxicology The peer-reviewed literature includes relatively little information on the toxicity of imidacloprid or other neonicotinoid insecticides in mammals. Research that has been published has generally dealt with a determination of acute lethal potency (e.g., LD50) for a series of structural analogues, without further assessment. One such study reported the presence of tremor in mice that had been treated with an acute oral dose of imidacloprid or one of several other neonicotinoids (Chao and Casida, 1997). This finding provides evidence of nicotinic stimulation at high (near-lethal or lethal) dose levels. A second source of information is provided by Yamamoto and Casida (1999),

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with chapters that discuss mammalian toxicology data for imidacloprid (Thyssen and Machemer, 1999), nitenpyram (Akayama and Minamida, 1999), and thiamethoxam (Maienfisch et al., 1999). Finally, there is a published comparison of the findings of neurotoxicity studies that were conducted in industry laboratories with commercial products (Sheets, 2001). This work is summarized in Section 95.11.4, following a review of the findings with imidacloprid. The general absence of published information on the toxicology of imidacloprid and other neonicotinoids in mammals contrasts with the extensive database generated by industry laboratories to support the registration of commercial products. The remainder of this chapter is largely devoted to a review of the toxicology studies that constitute the database that Bayer CropScience has generated for imidacloprid. These studies were conducted in accordance with regulatory guidelines, including those of the U.S. EPA (FIFRA), the OECD, and the Japanese MAFF, and in compliance with the associated Good Laboratory Practice (GLP) standards. The test material in these studies was technical-grade imidacloprid (purity of 94–98% active ingredient), with doses verified by analytical methods. For extended periods of exposure, imidacloprid was generally mixed in the diet and provided for ad libitum consumption, to model potential sources of human exposure and therefore to provide data that are best suited for risk assessments. Animals in these studies were acquired from commercial vendors as purpose-bred animals and were housed under standardized conditions that meet or exceed accepted standards for animal care. An overview of the results for acute toxicity, mutagenicity studies conducted with imidacloprid is provided in Tables 95.1 and 95.2 respectively.

95.5  Acute toxicity Imidacloprid was determined to produce minimal evidence of toxicity by acute dermal and inhalation routes of exposure and moderate acute toxicity by acute oral administration (Table 95.1). Imidacloprid is not an irritant and does not produce evidence of dermal sensitization. To assess acute oral toxicity, technical-grade imidacloprid was administered as an aqueous suspension to fasted, young-adult Wistar rats (five/sex/dose). Doses of 50 mg/kg in males and 100 mg/kg in females produced no evidence of toxicity. By comparison, higher doses of up to 315 mg/kg produced clinical signs in males and females, without causing mortality. At dose levels greater than 315 mg/kg, the incidence of mortality increased rather abruptly, with 20% mortality in both sexes at a dose of 400 mg/kg and 100% mortality at 500 mg/kg body weight. Clinical signs that were evident following treatment included tremor, gait incoordination, and evidence of decreased motility and activity, as

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Table 95.1  Acute Toxicity Studies with Technical-Grade Imidacloprida

well as nasal and urine staining. Signs of intoxication were evident within 15–40 min following oral administration and, with the exception of certain stains, were reversible within eight to 24 h following treatment. This outcome is consistent with the rapid distribution and metabolism profile that was summarized in Section 95.3. Treatment-related deaths generally occurred within 3–7 h following treatment.

Animal species

Route of exposure

LD50/LC50 (mg/kg BW/mg/m3 air)

Mouse

Oral

131–168

Rat

Oral

424–475

Rat

Dermal

5000

Rat

Inhalation aerosol 4 h

69b

95.6  Subchronic toxicity

Rat

Inhalation dust 4 h

5323c

95.6.1  Rat

Rabbit

Dermal

Not an irritant

Rabbit

Eye

Not an irritant

Guinea pig

Dermal

Negative for sensitizationd

a

LD50 and LC50 values represent the results for both sexes. Aerodynamic droplet size 5 M; 100%; maximum technically attainable concentration c Aerodynamic particle size 5 M; 4–11%. d Magnusson and Kligman test. b

Imidacloprid was administered through the diet for a period of 13 weeks to young-adult Wistar rats (10/sex/ dietary level) to examine cumulative toxicity with sustained exposure and to establish dietary levels suitable for the chronic toxicity/carcinogenicity study. In this study, the test substance was provided for ad libitum consumption at concentrations of 0, 150, 600, or 2400 ppm, which corresponded to average daily doses of 0, 14, 61, or 300 mg/kg body weight for males and 0, 20, 83, or 422 mg/kg body

Table 95.2  Mutagenicity Studies with Technical-Grade Imidacloprid Point mutation Salmonella microsome (AMES) test

Negative

Reverse mutation test E. coli

Negative

HPRT Chinese hamster ovary (CHO)

Negative

Chromosomal aberration in vitro Cytogenetics human lymphocytes

Slightly positive (at cytotoxic concentrations only)

Sister chromatid exchange (SCE) Chinese hamster ovary (CHO)

Slightly positive (at cytotoxic concentrations only)

Chromosomal aberration in vivo Micronucleus mouse bone marrow

Negative

Sister chromatid exchange Chinese hamster bone marrow

Negative

Cytogenetics Chinese hamster bone marrow

Negative

Cytogenetics mouse spermatogonia

Negative

Other genotoxicity tests Mitotic recombination yeast

Negative

Rec assay (B. subtilis)

Negative

Unscheduled DNA synthesis Rat hepatocytes

Negative

Chapter | 95  Imidacloprid: A Neonicotinoid Insecticide

weight for females. Satellite groups of control and highdose animals (10/sex/level) were retained for 4 weeks after the 13-week period of exposure to assess reversibility. Measures of brain and erythrocyte acetylcholinesterase and plasma cholinesterase activities were included to verify the expected absence of inhibition. Clinical signs associated with treatment were not evident in males or females at any of these dietary levels. Body weight and food consumption were reduced at the 600 ppm (males only) and 2400 ppm (both sexes) dietary levels. The average body weight for high-dose males and females was approximately 15% less than control. The liver was the principal target organ, with hypertrophy of hepatocytes and sporadic cell necrosis in high-dose males only. Liver pathology was mild at exposure termination and was fully reversible within the recovery period. Other effects in high-dose males and females included elevated serum alkaline phosphatase and alanine aminotransferase (ALAT) activities and a slight increase in blood clotting time. There was no inhibition of cholinesterase activity at any dietary level. The NOAEL (no-observed-adverse-effect level) for this study was 14 mg/kg/day in males and 83 mg/ kg/day in females.

95.6.2  Dog Toxicity was examined in a nonrodent species by administering imidacloprid through the diet for a period of 13 weeks to young-adult, pure-bred beagle dogs (four/ sex/level) at dietary concentrations of 0, 200, 600, or 1800/1200 ppm. These dietary exposure levels corresponded to average daily doses of 0, 8, 24, or 45 mg/kg body weight for males and females combined. Upon initiation of exposure, the 1800 ppm dietary level produced a sharp reduction in weight gain, relative to controls. Body weight regained after week 4, upon reducing the high dose to 1200 ppm. Tremor was evident in males and females in the 1800/1200 ppm dietary groups. However, tremor was not observed at comparable dietary levels in other studies using beagle dogs, including the 1-year dietary study (see Section 95.7.4). There was no evidence of tissue damage by clinical chemistry, gross necropsy examination, tissue weight, or microscopic examination at any dietary level. The NOAEL for this study was 24 mg/kg/day in both sexes.

95.7  Chronic toxicity and carcinogenicity 95.7.1  Rat A combined chronic toxicity/carcinogenicity study was conducted, with imidacloprid administered through the diet for a period of 2 years to Wistar rats (50/sex/dietary level).

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An additional set of animals (10/sex/level) was reserved for interim examination after 12 months. The test substance was provided for ad libitum consumption at concentrations of 0, 100, 300, 900, or 1800 ppm, which corresponded to average daily doses of 0, 5.7, 17, 51, or 103 mg/kg body weight for males and 0, 7.6, 25, 73, or 144 mg/kg body weight for females. Treatment-related clinical signs were not evident in either sex, and there was no effect on survival at any dietary level. Body weight was reduced by a maximum 12% in both sexes at the 1800 ppm dietary level and by 5–8% in males and females at the 900 ppm dietary level, but was not affected at lower levels of exposure. At 1800 ppm, serum alkaline phosphatase, creatine kinase, and aspartate aminotransferase (ASAT) activities were elevated and cholesterol was reduced. Microscopic lesions were also apparent in the thyroid at this dietary level, with mineralization of the colloid, fewer colloid aggregation sites, and parafollicular hyperplasia sites. These lesions were ascribed to an enhancement of biological aging processes and were not accompanied by a change in thyroid function (e.g., plasma T3, T4, and TSH levels were normal). Mineralization of thyroid colloid was also evident in males at 300 ppm and in both sexes at 900 ppm. There was no change in liver morphology and no inhibition of cholinesterase activity (brain, plasma, or erythrocyte) at any level. The NOAEL in this study was 5.7 mg/kg/day. There was no evidence of carcinogenicity at any dietary level.

95.7.2  Mouse To further assess oncogenic potential, imidacloprid was administered through the diet for a period of 24 months to B6C3F1 mice (50/sex/dietary level), at concentrations of 0, 100, 330, 1000, or 2000 ppm. An additional set of animals (10/sex/level) was reserved for interim examination after 12 months. These dietary concentrations resulted in average daily doses of 0, 20, 66, 208, or 414 mg/kg body weight for males and 0, 30, 104, 274, or 424 mg/kg body weight for females. There were no clinical signs associated with treatment and no effect on survival at any dietary level. Males and females that received the 2000 ppm dietary level had a marked decrease in body weight gain, relative to controls, with correspondingly lower food and water consumption. The difference in body weight reached 29% less than controls, indicating that this level exceeded a maximum tolerated dose (MTD). Liver changes were also evident at 2000 ppm but not at lower dietary levels. These consisted of low-grade periacinar hepatocyte hypertrophy, which was considered to represent metabolic adaptation to this xenobiotic. Effects that were evident at 1000 ppm consisted of reduced food consumption (females only) and reduced body weight, relative to controls, for males and females (up to 10% and 5%, respectively). There were no changes in serum chemistry, tissue weight, or tissue morphology

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(by gross and microscopic examination) associated with treatment at any dietary level. The number, type, distribution, and time of occurrence of tumors provided no evidence that imidacloprid has carcinogenic potential. The NOAEL in this study was 66 mg/kg/day.

95.7.3  Classification for Carcinogenicity Based on the collective results of the chronic toxicity and carcinogenicity studies in the rat and mouse, the U.S. EPA has classified imidacloprid in category E, indicating no evidence of carcinogenicity. There is no cancer risk associated with exposure to imidacloprid. This classification indicates that the database for imidacloprid supports evidence of noncarcinogenicity for humans. The reference dose (RfD) based on the 2-year rat feeding/ carcinogenicity study, with an NOAEL of 5.7 mg/kg body weight and 100-fold Uncertainty Factor is 0.057 mg/kg body weight. The theoretical maximum residue contribution (TMRC) from published uses is 0.008187 mg/kg body weight/day, utilizing 14.4% of the RfD.

95.7.4  Dog To evaluate chronic toxicity in a nonrodent species, imida­ cloprid was administered through the diet for a period of 52 weeks to young-adult, pure-bred beagle dogs (four/sex/ dietary level). The test substance was provided for ad libitum consumption at concentrations of 0, 200, 500, or 1250 ppm. The 1250 ppm dietary concentration was increased to 2500 ppm from week 17 onwards. These levels corresponded to daily doses of 0, 6.1, 15, and 41/72 mg/kg/day. The 1250 ppm dietary level was associated with a slight, but transient, fall in food consumption in both sexes with the introduction of treated feed. A similarly transient effect was evident when the dietary concentration was increased to 2500 ppm during week 17. The tremor that was reported in the subchronic dog study, at dietary concentrations of 1200/1800 (see Section 95.6.2), was not evident here at any dietary concentration. Effects at 1250 ppm included a slight increase in plasma cholesterol (females only) and a slight increase in hepatic cytochrome P450 activity (both sexes). The induction of cytochrome P450 enzymes was associated with a slight increase in liver weight. Thus, the liver was the principal target organ, but the respective changes represent metabolic adaptation to this xenobiotic. The chronic NOAEL in the dog was 72 mg/kg/day.

95.8  MUTAGENICITY Imidacloprid has been evaluated for mutagenicity using a full complement of in vitro and in vivo tests that is required for registration. These results indicate that imidacloprid is not mutagenic (Table 95.2). Briefly, the in vitro point

mutation tests were negative. This includes the results of chromosomal aberration tests conducted in vitro, which were negative at noncytotoxic concentrations and showed only slightly positive effects at cytotoxic concentrations. In vivo chromosomal aberration tests were also all negative. Finally, the mitotic recombination test that is conducted in yeast, the rec assay with Bacillus subtilis, and the unscheduled DNA synthesis (UDS) test were also all negative.

95.9  Developmental toxicity 95.9.1  Rat The potential for imidacloprid to produce developmental toxicity was examined in the rat. In this study, mated female Wistar rats (25/dose level) were treated daily, by gavage, on gestation days 6–15, with doses of 0, 10, 30, or 100 mg/kg body weight per day. On day 21 postcoitum, the fetuses were delivered by cesarean section and examined for development, including skeletal alterations. Dosages of 30 and 100 mg/kg/day produced signs of maternal toxicity (reduced food consumption and body weight gain). The highest dosage of 100 mg/kg/day produced a delay in fetal development and the offspring had wavy ribs as a reversible finding. While an increased incidence of wavy ribs, relative to controls, was ascribed to treatment, the incidence was within the range of historical controls. No fetal malformations were evident at any dose level. The maternal NOAEL was 10 mg/kg body weight per day and the fetal NOAEL was 30 mg/kg body weight per day. These results indicate that imidacloprid is not a primary embryotoxicant and is not teratogenic.

95.9.2  Rabbit The potential for imidacloprid to produce developmental toxicity was also examined in the rabbit. In this study, mated Chinchilla rabbits (16/dose level) were treated by gavage on gestation days 6 through 18, with daily doses of 0, 8, 24, or 72 mg/kg body weight. Cesarean section and examination of embryo and fetal development, including fetal skeletal alterations, were conducted on day 28 postcoitum. The highest dosage of 72 mg/kg/day produced severe maternal toxicity, including some deaths. Abortions and complete resorptions, delayed ossification, and reduced fetal weights were also evident at this dose level. The next lower dosage of 24 mg/kg/day produced slightly decreased food consumption and body weight gain, relative to controls, but no effects on the fetus. Thus, embryotoxicity was only evident at a maternally toxic dose. As with the rat, no fetal malformations were evident at any dose level. The maternal NOAEL was 8 mg/kg body weight per day and the fetal NOAEL was 24 mg/kg body weight per day. These results indicate that imidacloprid is not a primary embryotoxicant and is not teratogenic.

Chapter | 95  Imidacloprid: A Neonicotinoid Insecticide

95.10  Reproductive toxicity The potential effects of imidacloprid on reproduction and development were examined in a two-generation, twolitter study in Wistar rats (30/sex/dietary level in the parental generation). In this study, technical-grade imidacloprid was mixed in the diet for ad libitum consumption, at dietary concentrations of 0, 100, 250, and 700 ppm. The treated feed was provided during a prepairing period of 84 days and throughout pairing, gestation, and lactation for breeding of the F1A and F1B litters. Following weaning of the F1B litters on day 21 postpartum, the F1-generation parental animals were selected. The diets were fed for 105 days prior to pairing and throughout pairing, gestation, and lactation periods for breeding of the F2A and F2B litters. Gonadal function, estrus cycle, mating behavior, conception, parturition, lactation, weaning, and the growth and development of the offspring of multiple generations, as well as neonatal morbidity, mortality, and behavior were evaluated. For the parental generation, maternal toxicity was evident at the high dose as a decrease in body weight gain and food consumption, relative to controls. A pronounced reduction in body weight gain and food consumption, relative to controls, occurred during lactation. These effects coincided with the large increase in dietary exposure that occurs during lactation, when food consumption increases rather dramatically to support the offspring. Hepatic cytochrome P450, O-demethylase, and N-demethylase were also induced to high levels of activity in high-dose maternal animals. In the offspring, toxicity was evident at the high dose as a marked decrease in body weight gain, relative to controls, before weaning on postnatal day 21. There were no effects on reproduction or development at any dietary level. More specifically, there was no effect of treatment on mating indices, fertility, gestation, conception, litter size, or mortality, at any dietary level. There was also no evidence of pathology, in the form of malformations, gross lesions, a change in tissue weight, or histopathology at any exposure level. The NOAEL in this study was approximately 20 mg/kg body weight per day for the adult and approximately 35 mg/kg body weight per day for the offspring.

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(FOB) and a computer-automated test (figure-eight maze) to measure spontaneous activity, including habituation. At term (on day 14 after the acute dose or during week 14 of dietary exposure), a subset of the animals (six/sex/dose level) was anesthetized and perfused using aldehyde fixative, with representative skeletal muscle and neural tissues collected for microscopic examination.

95.11.2  Acute Neurotoxicity To evaluate acute neurotoxicity, Sprague-Dawley rats (12/ sex/dose level) received a single oral dose of imidacloprid, administered by gavage as an aqueous suspension at analytically confirmed doses of 0, 42, 151, or 307 mg/kg body weight in males and 0, 20, 42, 151, or 307 mg/kg body weight in females. Animals were evaluated using the FOB and the figure-eight maze 1 week prior to treatment, again at the time of peak neurobehavioral signs, which was approximately 4 h following treatment, and on days 7 and 14 following treatment. There was no evidence of systemic toxicity at 42 mg/kg. Compound-related effects at 151 mg/kg included tremor (one female), a slight decrease in body temperature, and red nasal stain. The 307 mg/kg dose produced severe acute toxicity, including lethality (two males and eight females), with these deaths occurring within 4–24 h following treatment. At the 4-h observation period, tremor was apparent in all surviving high-dose animals and was more severe, relative to the next lower dose. Body temperature was also reduced an average 2.0°C and 5.5°C in high-dose males and females, respectively. Additional effects at this lethal dose included evidence of motor incoordination (e.g., incoordinated gait and impaired aerial righting), autonomic signs (e.g., perianal and urine stains), and CNS depression (e.g., minimal activity and a diminished response to stimuli). Clinical signs following acute exposure generally resolved in surviving animals within 8–24 h following treatment. Urine stain was the only effect that persisted for up to 4 days after treatment. All findings in the FOB and the figure-eight maze had resolved by day 7, which was the first test occasion following the day 0 time-point. Neuropathology was not evident at the highest dose level. The NOAEL in this study was 42 mg/kg.

95.11  Neurotoxicity 95.11.1  General

95.11.3  Subchronic Neurotoxicity

The toxicology database for imidacloprid includes acute and subchronic neurotoxicity studies that were conducted in accordance with the U.S. EPA (FIFRA) guidelines. These studies were conducted using young-adult male and female rats, following an acute oral (gavage) dose or with 13 weeks of dietary exposure. In both studies, the animals were tested using a functional observational battery

To assess neurotoxicity with a sustained exposure, imidacloprid was administered via the diet for 13 weeks to young-adult Fischer-344 rats (12/sex/dietary level), at dietary concentrations of 0, 150, 1000, and 3000 ppm. These dietary levels resulted in average daily exposures of 0, 9.3, 63, and 196 mg/kg for males and 0, 10.5, 69, and 213 mg/kg for females. Neurobehavioral tests (FOB and

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activity in figure-eight maze) were performed 1 week prior to the initiation of treatment and during weeks 4, 8, and 13 of exposure. There was little evidence of toxicity in this study at any dietary concentration. Effects at dietary levels of 1000 ppm and 3000 ppm were generally limited to decreased food consumption and an associated decrease in body weight gain, relative to controls. The difference in body weight for high-dose males and females averaged 15% and 8% less than control, respectively. Toxicity was not evident by cage-side observation or the automated test of motor activity at any dietary level. On the last test occasion (week 13), there was a modest increase in the incidence of high-dose males with a slightly uncoordinated righting response that was ascribed to treatment. There was no evidence of neuro­ pathology at the high dose. The NOAEL for this study was 9.3 mg/kg body weight per day.

95.11.4  Comparison with other Neonicotinoids The results from the neurotoxicity studies with imidacloprid compare closely with the findings of the acute and subchronic neurotoxicity studies that were conducted in industry laboratories with acetamiprid, clothianidin, thiacloprid, and thiamethoxam (Sheets, 2001). Comparisons involving clothianidin and thiacloprid are facilitated by the fact that those studies were conducted under comparable conditions and in the same laboratory as the studies with imidacloprid. For acute neurotoxicity, the time of peak effects for these compounds ranged from 2 to 6 h following administration by gavage. The most consistent finding at lower doses was decreased activity, which was evident by observation and in the automated test devices. By comparison, the most common effects at higher dose levels were tremor, impaired pupillary function (either dilated or pin-point pupils), incoordinated gait, and hypothermia. In studies that included a lethal dose, deaths occurred within 4–24 h following treatment. Except for some residual staining, recovery generally occurred within 8–24 h following treatment. Neuropathology was not evident with any of these compounds. The results from subchronic neurotoxicity studies with acetamiprid, clothianidin, thiacloprid, and thiamethoxam are also comparable with the findings with imidacloprid. Each compound produced minimal effects, with decreased body weight and food consumption at higher dose levels and little or no overt evidence of an effect on the nervous system. Typically, there were no clinical signs, FOB findings, or effects on spontaneous activity in the automated devices, and little evidence of cumulative toxicity at any dietary level. Finally, none of these compounds produced neuropathology at the highest dietary level.

Hayes’ Handbook of Pesticide Toxicology

95.11.5  Developmental Neurotoxicity There is little information in the published literature to assist in determining the potential for imidacloprid, or any other neonicotinoid insecticide, to affect the developing nervous system. While the results from the developmental toxicity and multigeneration reproduction toxicology studies with imidacloprid provide no indication of developmental neurotoxicity, these studies are relatively limited in such assessment. Evidence of developmental neurotoxicity has been reported in rats following a single intraperitoneal injection of imidacloprid (337 mg/kg, 0.75  LD50, in corn oil) during gestation (Abou-Donia et al., 2008). Findings in this study included evidence of persistent behavioral changes and increased GFAP activity in the brain. However, these results are considered preliminary and of uncertain relevance to humans, based on the small sample size (five rats at a single dose level) and a route of exposure that differs from human circumstances. A more rigorous assessment of effects on the developing nervous system is provided from a study with imidacloprid that was conducted according to the U.S. EPA guideline for a developmental neurotoxicity study (U.S. EPA, OPPTS 870.6300). This study design includes a complement of automated tests of cognition, auditory startle habituation, motor activity ontogeny, and neuropathology, with exposures (route and duration) that are relevant to human circumstances and suitable for regulatory purposes in risk assessments. In this study, technical-grade imidacloprid (94% purity) was administered via the diet to mated female Wistar rats (30/dietary level) at nominal concentrations of 0, 100, 250 and 750 ppm from gestation day 0 through lactation day 21. The offspring from 20 litters/dietary level (minimum) were evaluated using detailed clinical observations, developmental landmarks (surface righting, acoustic startle, eye opening, vaginal patency, and balanopreputial separation), body weight, food consumption, spontaneous activity (figure-eight maze), auditory startle habituation, cognitive function (passive avoidance and M-maze), and ophthalmology, with neural tissues collected on postnatal days 11 and 75 (5 days) (study termination) for microscopic and morphometric examination. There was no effect on reproduction in this study at any dietary level. The only treatment effect in P-generation females was reduced food consumption at the high dose during gestation and lactation, with an associated decrease in body weight gain. Effects in the F1 generation were also limited to the high dose, with decreased weight gain, such that body weight averaged 11–13% less than control on postnatal days 4–21. Also at the high dose, measures of activity were less than control on postnatal day 17 (males and females) and postnatal day 21 (females only). Developmental landmarks and other neurobehavioral tests were not affected at any dietary level. A modest decrease in thickness of the caudate putamen in high-dose females on postnatal day 75, relative to control, was considered

Chapter | 95  Imidacloprid: A Neonicotinoid Insecticide

an effect by the U.S. EPA. However, this difference from the concurrent control group was not considered an effect due to treatment by the report author, because the results for the high-dose females were within the range of historical control and there was no similar difference for high-dose males or in either gender at the younger age. There was no effect on brain weight at any dietary level and no gross or microscopic lesions in high-dose males or females. The maternal and offspring NOAEL was 20 mg/kg/day.

Conclusion Imidacloprid [1-[(6-chloro-3-pyridiny1)methyl]-N-nitro-2imidazolidinimine] is the first representative of the neonicotinoid insecticides that was registered for use and remains the most important commercial product. The insecticidal activity of imidacloprid and other neonicotinoids is attributed to actions on post-synaptic nicotinic receptors, with relatively low toxicity in mammals due to differences in binding properties to the various nicotinic receptor subtypes in vertebrate species, as well as rapid elimination and poor penetration of the blood–brain barrier, Acute toxicity is characterized by nicotinic signs (e.g., tremor) at relatively high levels of exposure. Imidacloprid produces minimal evidence of toxicity by acute dermal and inhalation routes of exposure, moderate acute toxicity by oral administration, is not an irritant and does not produce evidence of dermal sensitization. Imidacloprid is not mutagenic or carcinogenic, it is not a primary embryotoxicant, is not teratogenic and has no effect on reproduction or development.

References Abou-Donia, M. B., Goldstein, L. B., Bullman, S., Tu, T., Khan1, W. A., Dechkovskaia1, A. M., and Abdel-Rahman, A. A. (2008). Imidacloprid induces neurobehavioral deficits and increases expression of glial fibrillary acidic protein in the motor cortex and hippocampus in offspring rats following in utero exposure. J. Toxicol. Environ. Health, Part A 71, 119–130. Akayama, A., and Minamida, I. (1999). Discovery of a new systemic insecticide, nitenpyram and its insecticidal properties. In “Nicotinoid Insecticides and the Nicotinic Acetylcholine Receptor” (I. Yamamoto and J. E. Casida, eds.), pp. 127–148. Springer-Verlag, Tokyo. Buckingham, S. D., Lapied, B., Le Corrone, H., Grolleau, F., and Sattelle, D. B. (1997). Imidacloprid actions on insect neuronal acetylcholine receptors. J. Exp. Biol. 200, 2685–2692. Chao, S. L., and Casida, J. E. (1997). Interaction of imidacloprid metabolites and analogs with the nicotinic acetylcholine receptor of mouse brain in relation to toxicity. Pest. Biochem. Physiol. 58, 77–88. Gundelfinger, E. D., and Schulz, R. (2000). Insect nicotinic acetylcholine receptors: Genes, structure, physiological and pharmacological properties. In “Handbook of Experimental Pharmacology, Vol. 144, Neuronal Nicotine Receptors” (F. Clementi, D. Fornasari, and C. Gotti, eds.), pp. 497–521. Springer-Verlag, Tokyo.

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Kagabu, S., Moriya, K., Shibuya, K., Hattori, Y., Tsuboi, S., and Shiokawa, K. (1992). 1-(6-Halonicotinyl)-2-nitromethylene-imidazolidines as potential new insecticides. Biosci. Biotech. Biochem. 56(2), 362–363. Leicht, W. (1993). Imidacloprid: A chloronicotinyl insecticide. Pestic. Outlook 4(3), 17–21. Liu, M.-Y., and Casida, J. E. (1993). High affinity binding of [3H]Imidacloprid in the insect acetylcholine receptor. Pestic. Biochem. Physiol. 46, 40–46. Liu, M.-Y., Lanford, J., and Casida, J. E. (1993). Relevance of [3H]Imidacloprid binding site in house fly head acetylcholine receptor to insecticidal activity of 2-nitromethylene- and 2-nitroimino-imidazolidines. Pestic. Biochem. Physiol. 46, 200–206. Maienfisch, P., Brandl, F., Kobel, W., Rindlisbacher, A., and Senn, R. (1999). CGA 293’343: A novel, broad-spectrum neonicotinoid insecticide. In “Nicotinoid Insecticides and the Nicotinic Acetylcholine Receptor” (I. Yamamoto and J. E. Casida, eds.), pp. 177–209. Springer-Verlag, Tokyo. Matsuda, K., Buckingham, S. D., Freeman, J. C., Squire, M. D., Baylis, H. A., and Sattelle, D. B. (1998). Effects of the alpha subunit on imidacloprid sensitivity of recombinant nicotinic acetylcholine receptors. Br. J. Pharmacol. 123, 518–524. Methfessel, C. (1992). Effect of imidacloprid on the nicotinergic acetylcholine receptors of rat muscle. Pflanzenschutz Nachricten Bayer 45, 369–380. Minamida, I., Iwanaga, K., Tabuchi, T., Aoki, I., Fusaka, T., Ishizuka, H., and Okauchi, T. (1993). Synthesis and insecticidal activity of acyclic nitroethene compounds containing a heteroarylmethylamino group. J. Pesticide Sci. 18, 41. Nagata, K., Song, J. H., Shono, T., and Narahashi, T. (1998). Modulation of the neuronal nicotinic acetylcholine receptor-channel by the nitromethylene heterocycle imidacloprid. J. Pharmacol. Exper. Ther. 285, 731–738. Nagata, K., Aoyama, E., Ikeda, T., and Shono, T. (1999). Effects of nitenpyram on the neuronal nicotinic acetylcholine receptor-channel in rat phaeochromocytoma PC12 cells. J. Pesticide Sci. 24, 143–148. Schroeder, M. E., and Flattum, R. F. (1984). The mode of action and neurotoxic properties of the nitromethylene heterocycle insecticides. Pest. Biochem. Physiol. 22, 148–160. Sheets, L. P. (2001). Neonicotinoid Insecticides. In “Neurotoxicology Handbook” (E. J. Massaro, ed.), Vol. 1. Humana Press, Totowa, NJ. Shiokawa, K., Tsuboi, S., Kagabu, S., and Moriya, K. (1986). Jpn. Kokai Tokkyo Koho JP 61–267575. Soloway, S. B., Henry, A. C., Kollmeyer, W. D., Padgett, W. M., Powell, J. E., Roman, S. A., Tieman, C. H., Corey, R. A., and Horne, C. A. (1978). Nitromethylene insecticides. Adv. Pestic. Sci. 4, 206–217. Takahashi, H., Mitsui, J., Takakusa, N., Matsuda, M., Yoneda, H., Suzuki, J., Ishimitsu, K., and Kishmoto, T. (1992). NI-25, a new type of systemic and broad spectrum insecticide. In “Brighton Crop Protection Conferences B Pest and Diseases,” Vol. 1, pp. 89–96. Thyssen, J., and Machemer, L. (1999). Imidacloprid: Toxicology and metabolism. In “Nicotinoid Insecticides and the Nicotinic Acetylcholine Receptor” (I. Yamamoto and J. E. Casida, eds.), pp. 213–222. Springer-Verlag, Tokyo. Tomizawa, M., and Casida, J. E. (1999). Minor structural changes in nicotinoid insecticides confer differential subtype selectivity for mammalian nicotinic acetylcholine receptors. Br. J. Pharmacol. 127, 115–122. Tomizawa, M., and Yamamoto, I. (1993). Structure-activity relationships of nicotinoids and imidacloprid analogs. J. Pesticide Sci. 18, 91–98. Tomizawa, M., Latli, B., and Casida, J. E. (1999). Structure and function of insect nicotinic acetylcholine receptors studied with nicotinoid insecticide affinity probes. In “Nicotinoid Insecticides and the Nicotinic Acetylcholine Receptor” (I. Yamamoto and J. E. Casida, eds.), pp. 271–292. Springer-Verlag, Tokyo.

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Wakita, T., Kinoshita, K., Yamada, E., Yasui, N., Kawahara, N., Naoi, A., Naka, M., Ebihara, K., Matsuno, H., and Kodaka, K. (2003). The discovery of dinotefuran: A novel neonicotinoid. Pest. Manag. Sci. 59, 1016–1022. Wiesner, P., and Kayser, H. (2000). Characterization of nicotinic acetylcholine receptors from the insects Aphis craccivora. Myzus persicae, and Locusta migratoria by radioligand binding assays: Relation to thiamethoxam action. J. Biochem. Mol. Toxicol. 14, 221–230. Yamada, T., Takashi, H., and Hatano, R. (1999). A novel insecticide, acetamiprid. In “Nicotinoid Insecticides and the Nicotinic Acetylcholine Receptor” (I. Yamamoto and J. E. Casida, eds.), pp. 149–176. Springer-Verlag, Tokyo. Yamamoto, I., and Casida, J. E. (1999). “Nicotinoid Insecticides and the Nicotinic Acetylcholine Receptor.” Springer-Verlag, Tokyo. Yamamoto, I., Yabuta, G., Tomizawa, M., Saito, T., Miyamoto, T., and Kagabu, S. (1995). Molecular mechanism for selective toxicity of nicotinoids and neonicotinoids. J. Pesticide Sci. 20, 33–40.

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Yamamoto, I., Tomizawa, M., Saito, T., Miyamoto, T., Walcott, E. C., and Sumikawa, K. (1998). Structural factors contributing to insecticidal and selective actions of neonicotinoids. Arch. Insect Biochem. Physiol. 37, 24–32. Zhang, A., Kayser, H., Maienfisch, P., and Casida, J. E. (2000). Insect nicotinic acetylcholine receptor: Conserved neonicotinoid specificity of [3H]Imidacloprid binding site. J. Neurochem. 75, 1294–1303. Zwart, R., Oortgiesen, M., and Vijverberg, H. P. M. (1992). The nitromethylene heterocycle 1-(pyridin-3-yl-methyl)-2-nitromethyleneimidazolidine distinguishes mammalian from insect nicotinic receptor subtypes. Eur. J. Pharmacol. 228, 165–169. Zwart, R., Oortgiesen, M., and Vijverberg, H. P. M. (1994). Nitro­ methylene heterocycles: Selective agonists of nicotinic receptors in locust neurons compared to mouse N1E-115 and BC3H1 cells. Pestic. Biochem. Physiol. 48, 202–213.

Chapter 96

Interactions with the Gamma-Aminobutyric Acid A-Receptor: Polychlorocycloalkanes and Recent Congeners and Other Ligands Gerald T. Brooks Burgess Hill, United Kingdom

96.1  Introduction Chlorinated insecticides have been with us for 60 years (see Plimmer et al., 2003). With the exception of lindane (gammahexachlorocyclohexane, HCH) and endosulfan, which are relatively biodegradable and still find extensive uses, most have already been phased out or are being phased out. Their insecticidal properties were discovered at a time when the study of biochemical toxicology was in its infancy. However, the metabolism of dichlorodiphenyl-trichloroethane (DDT) was soon discovered after resistance to it appeared in 1947 and the similarities between the actions of DDT and the natural pyrethrins and cross-resistance to them in insects provided the stimulus that soon led to recognition of DDT action on the sodium channel of nerve membrane. The history of lindane and the cyclodiene-related group (collectively polychloro­cycloalkanes, PCCAs) is more complex, partly because of the variety of commercially viable insecticides that arose from the early discoveries. Lindane was soon found to be biodegradable but the strongly residual nature of the cyclodiene insecticides and the discovery that aldrin was converted into its stable epoxide, dieldrin, led to the view that these insecticides were inert. Moreover, apart from lindane, there were no other insecticide classes with recognized similar toxic action at the time and the mode of action was completely unknown and would not be revealed until more than 30 years later (ca. 1982)! A comprehensive account of the salient toxicology of these compounds was given in the first edition of this handbook (Smith, 1991). The present account summarizes research on the cyclodiene and related insecticides, which led to an appreciation of their structure–toxicity relationships and in the end to an understanding of their mode of Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

action as noncompetitive antagonists acting in the chloride ion channel of the gamma-aminobutyric acid (GABA) Areceptor. Developments subsequent to this discovery make research in this area a subject of continuing fascination. For uniformity, the chemical nomenclature used follows that in the first edition. Other systems are in use: see Bedford (1974) and Brooks (1974). Simple acronyms and common names will be used wherever possible for chemical compounds as the full chemical names become cumbersome, especially for some of the skeletal rearrangement products so common in this series.

96.2  Discovery of polychlorocycloalkane metabolism as a factor in toxicity 96.2.1  Background According to the account of Lauger et al. (1944), DDT was the first molecule rationally designed as an insecticide, based on the known fumigant properties of chlorobenzene and the anesthetic properties of highly lipophilic chloroform. In contrast, the insecticidal properties of technical-HCH (t-HCH) (Bender, 1935) and the first cyclodiene insecticides (Hyman, 1949; Kearns et al., 1945) were discovered as a result of the commercial interest in new uses for readily available chlorine and for hydrocarbons such as benzene and cyclopentadiene, chlorinated hydrocarbons being of general interest, for example, as dielectrics and fire retardants. Thus, Bender added benzene to liquid chlorine in a field and noticed that the product killed insects. Hyman sought new uses for cyclopentadiene; hexachlorocyclopentadiene (“hex”) was known 2065

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to be stable and, at first surprisingly, was found to react easily with cyclopentadiene in a Diels-Alder reaction, which led to chlordene, and later with norbornadiene (NB) to give aldrin. The addition of two chlorines to chlordene gave the chlordane isomers, with greatly increased insecticidal potency, whereas allylic chlorination gave heptachlor. A variant of the synthesis of aldrin, in which hexachloronorbornadiene (HCNB) reacted with cyclopentadiene, gave the isomeric isodrin and both compounds underwent chemical epoxidation to their crystalline epoxides, dieldrin and endrin, respectively. These heavily chlorinated insecticides were at that time considered to be rather inert, whereas t-HCH was long known to be readily dechlorinated to trichlorobenzenes, etc., which was its practical use. The potent insecticidal activity of lindane (gamma-HCH; 1, Figure 96.1) was not established until 1943, 10 years after Bender’s original observation, because lindane comprised only 10–15% of t-HCH and was readily lost during purification, which resulted mainly in crystalline, but inactive alpha- and beta-HCH (Slade, 1945). It is unfortunate that due to the high potency of lindane, t-HCH could be used directly and extensively as a practical insecticide, resulting in contamination of the environment with the remaining inactive isomers; lindane itself is relatively biodegradable and continues to be a valuable insecticide. Cl

Cl

Cl

Cl

Cl6

Cl

Cl

Cl

Cl

Cl

Cl

O

Cl

(1) LINDANE

Cl (2)

Cl6

(3) Cl

Cl6

Cl6

All of these discoveries predate modern biochemical toxicology. Indeed, resistance to modern insecticides, beginning with DDT in 1947, afforded the initial stimulus for research in this area, which subsequently became known as insect, or insecticide, toxicology and developed in parallel with but somewhat behind mammalian toxicology. One major mechanism of insect resistance to DDT was eventually found to involve its enzymatic dehydrochlorination to DDE (Sternburg et al., 1954). When it was discovered that certain nontoxic DDT analogues and some other compounds suppressed resistance when co-applied with DDT, studies of the mechanisms of this synergistic effect became an important aspect of insect toxicology and synergists later became standard tools for the detection of metabolic detoxication. Natural pyrethrins were well known to be strongly synergized by various inactive methylenedioxyphenyl derivatives (e.g., piperonyl butoxide: PBO) but, as esters, these insecticides were considered likely to be hydrolyzed in vivo and the mechanism of the synergistic effect was not understood. Insect resistance to the cyclodienes became evident in the early 1950s, but from research conducted on housefly resistance after 1957 (Brooks, 1960), it appeared not to involve enzymatic detoxication, in contrast to the situation with DDT. Meanwhile, Ryan and Engel (1957) found that carbon monoxide inhibited the C21-hydroxylation of 17-hydroxyprogesterone by microsomes from the vertebrate adrenal cortex and showed this inhibition to be lightreversible; Klingenberg (1958) reported that rat liver microsomes contained a similar pigment, subsequently called cytochrome P450 (CyP450) (Estabrook et al., 1963), which appeared to be important in the metabolism of steroids and drugs, and, in 1965, this pigment was shown to be present in microsomal preparations from insects (Lewis, 1967; Ray, 1967). Forty years on, CyP450 research is a routine adjunct to studies on pesticide resistance and toxicology (Hodgson and Rose, 2008).

O

(4)

(5)

Cl6

(6)

Cl6 O

O

(7) HEOM

(8) HCE

Cl6

96.2.2  Lindane, Aldrin, Dieldrin, Isodrin, and Endrin and Analogues

Cl6

Cl6

(9) Cl6

O

(10)

(11)

(12)

Cl6 O O

(13) Figure 96.1  Chemical structures of compounds mentioned in the text.

A new age dawned in insect toxicology in 1960, when Sun and Johnson published the results of synergism experiments with several classes of organophosphorus insecticides and some cyclodienes (Sun and Johnson, 1960). The latter showed small factors of either antagonism (for aldrin, 2, Figure 96.1) or synergism (for dieldrin, 3; isodrin, 4; and endrin, 5) when used in combination with the methylenedioxyphenyl (MDP) synergist sesoxane (sesamex), representative of the well-known pyrethrin synergist structures. They suggested that their results could be explained by the inhibition of metabolic oxidations in vivo, showed that sesoxane inhibited the epoxidation of aldrin in vivo, and postulated that the long-known synergism of pyrethrins by methylenedioxyphenyl compounds resulted, in fact, from inhibition of their

Chapter | 96  Interactions with the Gamma-Aminobutyric Acid A-Receptor

oxidative detoxication. The small factors of antagonism for aldrin and heptachlor (6, Figure 96.1) suggested that epoxidation, their only reported biotransformation at that time, was a bioactivation (toxication) reaction, so that the precursors were possibly propesticides (in current terminology). At this time, numerous nonepoxide cyclodiene analogues were found to be synergized by sesoxane (Brooks and Harrison, 1963, 1964a), indicating that they had intrinsic toxicity of their own, although they may be pharmacokinetically less efficient than the epoxides. Also, epoxidation generally produces another toxicant, so that the level of toxic material in the tissues is maintained, whereas it is attenuated if the conversion is a detoxication reaction. Remarkably, considering the small structural change involved, removal of the unchlorinated methanobridge from dieldrin gave the isomeric cyclohexane-derived epoxides HEOM and HCE (7 and 8, Figure 96.1), which were inactive (HEOM) and poorly toxic (HCE). With sesoxane, however, HCE became as toxic as dieldrin to houseflies, whereas HEOM toxicity was not improved. HCE was then found to be hydroxylated by mixed-function oxidases (MFOs) in vivo, mostly with epoxide ring retention, whereas HEOM suffered only addition of water to the epoxide ring, which was found to be an enzymic process, not inhibited by sesoxane (Brooks, 1966). The relative efficiencies of these pathways (Figure 96.2) in vitro are shown in Table 96.1, from which it is evident that, for HCE, oxidation is the main route in microsomes from several species; liver microsomes from birds and the rat hydroxylated HEOM to some extent, whereas those from rabbit and pig liver and houseflies hydrated the epoxide ring too rapidly for oxidation to be observed. Similar products and their conjugates are formed in vivo (Chipman and Walker, 1979), and the availability of two routes, one blockable by MFO-inhibiting synergists in vivo in insects, offered the possibility of selective toxicity in favor of vertebrates, which have epoxide ring hydration available as an escape route. Because the hydration of HEOM could not be

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significantly inhibited in vivo, its intrinsic toxicity was not demonstrable by the use of any known synergist. However, HEOM had the same toxicity as DDT to tsetse flies and was toxic to some species of mosquitoes, which appeared not to hydrate the epoxide ring efficiently, thus establishing that HEOM was intrinsically toxic (Brooks et al., 1981).

O OH

A

OH INHIBITORS OF MICROSOMAL OXIDATION NADPH, BLOCK R, B O2

O

NADPH, O2

O Cl6

Cl6 HEOM (3)

(1) HCE RA, P, B, H2O

OH

O

H, P, R, EPOXIDE HYDROLASE RA NOT READILY BLOCKED BY INHIBITORS

H2O

OH (2)

(4)

OH

OH

Figure 96.2  Alternative phase I metabolic pathways for HCE and HEOM and the effect of inhibitors. Epoxide ring hydration (1 gives 2 and 3 gives 4) is enzymic and only one enantiomer of HCE is hydrated: A, all species examined; B, some birds; H, housefly; P, pig; R, rat; Ra, rabbit. Pathways in vitro and in vivo for housefly; for mammals and birds, results are for liver homogenates and microsomes, but similar products and their conjugates are formed in vivo.

Table 96.1  Oxidation vs. Hydration of HCE and Hydration of HEOM by Microsomes from Vertebrate Liver and Houseflies Pigeon

Quail

Rat

Housefly

Rabbit

Pig

1.0

0.7c

1.3

1.3

2.0

1.0

0.02

0.03

0

0.33

1.0

0.06

5.74

37.0

46.0

100

a

HCE

  Oxidationb b

  Hydration

0

c

d

HEOM

  Hydration ratee 0.005 a

Vertebrate liver and housefly abdomen microsomes (NADPH); incubations for birds 90 min at 42°C; others 30 min at 37°C. Conversion of epoxide (HCE), percent/min at pH 7.4. c 11,000 g supernatant. d Incubations with pig liver and housefly microsomes at 30°C for 30 min (pH 8.4), with rabbit and rat liver microsomes at 37°C (pH 7.4), and bird liver microsomes (El Zorgani et al., 1970) at 42°C (pH 7.4). e Relative to pig liver (100 is equivalent to 31 g HEOM-diol formed/mg microsomal protein/min). b

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These observations verified the Sun-Johnson hypothesis regarding the action of MDP synergists, completely altered the perspective regarding the “inertness” of cyclodienes, and provided the first firm evidence for the existence of epoxide hydrolases (EHs), which Boyland (1950) had suggested to mediate the ultimate metabolism of aromatic hydrocarbons via labile, nonisolable epoxides (see Morisseau and Hammock, 2008). Also, sesoxane was found to stabilize certain of the metabolically labile cyclodienes in both dieldrin-resistant (R-) and dieldrin-susceptible (S-) houseflies but it synergized them only in S-flies, supporting the view that resistance did not involve metabolic detoxication (Brooks and Harrison, 1964b). Korte and Arent (1965) reported that dieldrin-treated rabbits excreted trans-6,7-dihydroxy-6,7-dihydroaldrin (t-DDA; 1, Figure 96.3) in their urine, indicating the epoxide ring opening of dieldrin to occur in vivo. Dieldrin and the heptachlor epoxides were subsequently found to be hydrated slowly in microsomal preparations from the livers of rabbits and pigs (Brooks and Harrison, 1969b; Brooks et al., 1970).

COOH

CONJUGATION

COOH R (V) OH

trans-

Cl mice (v)

PHOTODIELDRIN

+ CONJUGATES (v) (i) OH (1)

Ra (v) (i)

(2)

cis-

Cl

environment: rearrangement Cl Cl

Cl

Cl

mo (v) H (v) R (v) mfo

S

I (v)

hydration

H

H

OH

Cl

O

Cl

Cl

OH

P(i) S(v) R(v) (i) I(v), (m)

Cl

Cl

Cl

O Cl

Cl

Cl

man (v) R (v) (i) S (v)

mfo

mfo

S

R (v) (i) mice (v)

Cl

Cl

mice (v)

O

Cl

Cl

H

+ H Cl OH

Cl

O H

Cl

Cl

OH

Cl

HO ENDRIN Cl

O

H

H

Cl Cl

R (v)

O (7) H

O

OH

mfo OH

‘KLEIN KETONE’

O

mfo

Cl

H

(3)

(5)

mfo R (i)

O

Cl

O-GLUCURONIDE

OH

H H

Cl transDIOL

O H

O

Cl

? Cl

Cl

(8)

?

Cl

Cl

R (i) Cl (4)

H

OH

Cl

mfo

O

Cl

Cl

DIELDRIN

Cl

O

(6)

O O

Figure 96.3  Biotransformation routes of dieldrin, photodieldrin, and endrin. Wavy line indicates only partial ring structure shown. H, housefly; I, some insects; mo, mosquito; m, microorganisms; P, pig; Ra, rabbit; R, rat; S, sheep; v, in vivo; i, in vitro; S →, sesoxane inhibits. See text for references.

This challenged the hitherto prevailing view that cyclodienes would accumulate indefinitely in the tissues of treated animals, a challenge reinforced by the pharmacokinetic studies of Ludwig et al. (1964) and Robinson et al. [cited in Brooks (1969)] on mammals, birds, and marine organisms. The continual improvement in techniques for preparing microsomes from liver and other animal tissues afforded opportunities for the rapid examination of the likely phase 1 metabolites and stimulated interspecies comparisons of cyclodiene metabolism (Craven et al., 1976; El Zorgani et al., 1970; Slade et al., 1975). In contrast to alpha- and especially beta-HCH, lindane proved to be quite biodegradable, and its complex metabolism in insects and vertebrates via dehydrochlorination and oxidation to chlorinated phenols and their conjugates is well documented (Brooks, 1974; Smith, 1991; Ullman, 1972). In contrast to the conversion of aldrin and isodrin into their stable 6,7-epoxides, dieldrin and endrin (Giannotti et al., 1957; Kunze and Laug, 1953), 6,7-dihydroaldrin (9, Figure 96.1) and 6,7-dihydroisodrin (10) lacking the olefinic double bond, are monohydroxylated in the 6-(7) position by microsomal oxidases (Brooks, 1966; Brooks and Harrison, 1969a) and in vivo in houseflies. The stereo­ selective monohydroxylation of dihydroisodrin provided an alternative to the epoxidation reaction for the measurement of MFO activity (Krieger, 2008). The synergism of these dihydro-compounds by sesoxane in this insect indicates that this is a detoxication, in contrast to the epoxidation reactions. This result also suggests that the dihydro-compounds are intrinsically toxic, although they act more slowly than the epoxides. Notably, these dihydro-compounds have the same synergized toxi­ city as photoisodrin (11), the complete cage rearrangement product of isodrin, which is as polar as endrin (based on Rf values) and acts more rapidly than the dihydro-compounds. Also, the 5,8-oxirane (12), which has a “built-in” epoxide function, is similar in toxicity to dieldrin when synergized and acts more rapidly than the other dihydrocompounds discussed (Brooks, 1966; Brooks and Harrison, 1963), again supporting the view that such compounds are intrinsically active and that epoxidation (or an appropriate increase in polarity) improves the pharmacokinetic properties of these molecules, besides maintaining the total level of toxicant in the tissues. This does not, however, rule out a possibly more efficient binding of the epoxides at the site of action. Insect poisoning by commercial cyclodienes was not at first recognized to be reversible because their persistence in the tissues led to eventual death due to desiccation and starvation, without any recovery. Housefly poisoning by HCE was noted to be reversible, however, and even insects poisoned with HCE/sesoxane combinations would occasionally recover after prolonged periods of knockdown, although their wing musculature appeared to be permanently damaged. A further complication arose when

Chapter | 96  Interactions with the Gamma-Aminobutyric Acid A-Receptor

6,7-dihydroxydihydroaldrin (aldrin-trans-diol; t-DDA; 1, Figure 96.3) (Wang et al., 1971) and subsequently the corresponding cis-DDA (Burt, 1973) were found to be rapidly neuroactive when applied to isolated nerve ganglia of the American cockroach (Periplaneta americana), in contrast to dieldrin, which acted significantly more slowly. Wang et al. (1971) then suggested that the slow action of dieldrin was related to a requirement for its conversion into transDDA, as the active neurotoxicant liberated at the site of action by dieldrin hydration. Small amounts of these diols were later reported to be dieldrin metabolites in this insect (Nelson and Matsumura, 1973). However, t-DDA caused prostration only slowly when injected into cockroaches, quite different from the rapid action of dieldrin in vivo. The bioactivation hypothesis was further supported by the neuroactivity of t-DDA observed on frog nerve-muscle preparations (Akkermans et al., 1974, 1975a,b). The findings for cockroach were confirmed (Schroeder et al., 1977; Shankland and Schroeder, 1973), but, based on the less intense neuroactive effect of the diols and their very slow action in vivo, the diols were concluded to be detoxification products in the cockroach. Both diols are produced as metabolites of dieldrin by rats and mice and appear to be detoxification products in these mammals. These pharmacokinetic studies raise questions about possible internal barriers to the penetration of such molecules and their metabolites to critical sites in the nervous system. Similar problems are apparent throughout the series and are difficult to resolve experimentally. Moreover, a particular metabolite might be a bioactivation product in one species but a detoxification product in another. The tendency for molecular rearrangements in the environment (e.g., from exposure to sunlight) and in vivo has complicated investigations on residues and metabol­ites. Photoconversion products are frequently more toxic than the parent insecticides and may themselves be further metabolized; for example, photodieldrin (PD; 2, Figure 96.3) is oxidatively dechlorinated to the pentachloroketone (3, Figure 96.3; “Klein’s ketone”; Klein et al., 1970) in rats and insects (Baldwin and Robinson, 1969; Baldwin et al., 1972; Khan et al., 1970; Matthews and Matsumura, 1969); PD has a much shorter half-life (2–3 days) than dieldrin (10–13 days) in rat adipose tissue but is two- to fourfold more toxic to rodents and insects (Table 96.2). Dieldrintreated rats excrete 9-hydroxy-dieldrin (9-HD; 4, Figure 96.3) in the feces and the pentachloroketone in the urine (Richardson et al., 1968), and these are considered to arise by alternative modes of attack from beneath the ring system (Figure 96.3). The same pentachloroketone (3) was produced, along with varying amounts of 9-HD and cisand trans-DDA, in American cockroaches, German cockroaches (Blattella germanica), and houseflies (Nelson and Matsumura, 1973). The pentachloroketone (3) was reported to be more toxic than photodieldrin to mosquitoes and houseflies (Khan et al.,

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Table 96.2  Toxicity Data for Some Polychlorocyclohexane Insecticides and their Transformation Products Compound

Rodent acute oral Topical 24-h LD50a (mg/kg) LD50b housefly (g/g)

Lindane

90–190

1.5

Aldrin

38–60

1.5

6,7-Dihydroaldrin

40 (3.0)

5,8Oxadihydroaldrin

30 (1.5)

Dieldrin

47

1.0

77 (m) HCE

400

90 (2.0)

200–400 (m) HEOM Photodieldrin

500 10

0.12

7 (m) Didechlorodieldrin (DD)

0.9

0.2

1.4 9-Hydroxydieldrin (9-HD)

400 (m)

750

trans-Dihydroaldrin- 1,250 (m) diol (t-DDA)

750

Isodrin

3.0

12–17

6,7-Dihydroisodrin

39 (4.0)

Photoisodrin

2,000

15 (3.0)

Endrin

5.6

2.0

29 (m) 9-Keto-endrin

1.0

0.95

anti-9-Hydroxyendrin (AHEN)

2.5–5.5

100

syn-9-Hydroxyendrin

1.2

1.2

Heptachlor epoxide (HE160)

60

1.0

Heptachlor epoxide (HE90)

6.0

1-Hydroxychlordene 2,400–4,600

Inactive

Chlordene

50 (20)

Chlordene exoepoxide

35 (4.0) (Continued)

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Table 96.2  (Continued) Compound

Rodent acute oral Topical 24-h LD50a (mg/kg) LD50b housefly (g/g)

1-Hydroxy-chlordene exo-epoxide

Inactive

trans-Chlordane

1,100

11

cis-Chlordane

500–600

4.0

alpha-Endosulfan

76

5.0

beta-Endosulfan

240

9.0

Endosulfan sulfate

76

9.5

Endosulfan diol

15,000

500

Endosulfan ether

15,000

500

alpha-Hydroxyendosulfan ether

1,750

500

Endosulfan lactone

306 (m)

500

Isobenzan

3–10

1.0

6.0 (m) Bromocyclen (Bromodan®)

13,000

11.5

Chlorbicyclen (Alodan®)

15,000

15.5

Mirex

600–  3,000

Chlordecone

125

Toxaphene (technical)

90–270

Toxaphene (component B)

75 (m; ip)

Data from Bedford et al. (1975a); Brooks and Harrison (1964a); Buchel et al. (1966a, b); Jager (1970); Khan et al. (1970); Korte (1967); MaierBode (1968); Miles et al. (1969); Smith (1991). a For rat unless marked (m) for mouse. b Parenthetic values in housefly column are toxicities measured with sesoxane (5 g), preapplied before the insecticide to inhibit microsomal oxidases.

1970) but less toxic and slower acting than PD to the German cockroach (Kadous and Matsumura, 1982; Reddy and Khan, 1977), indicating that PD itself is the active toxicant in this insect. PD acted fourfold more rapidly (LD50, 0.01 g/insect) than dieldrin (LD50, 0.05 g/insect) and twofold more rapidly than the pentachloroketone (LD50, 0.13 g/insect), observations that suggest it has pharmacokinetic properties more favorable for toxicity than the other compounds. 9-HD (4) appeared to be more toxic (LD50, 0.02 g/insect) than dieldrin to the German cockroach and may contribute to dieldrin’s toxicity in this insect; the cisand trans-DDAs appeared to be relatively nontoxic when

injected. From other experiments on the American cockroach, it seems clear that these metabolites can enter the nerve cord from the insect body and are also produced in small amounts by metabolism in the nerve itself. Isodrin was found to be epoxidized to endrin (Figure 96.3) in houseflies (Brooks, 1960) and subsequently by liver microsomes from rats and rabbits, as a result of mixed-function oxidase (MFO) action (Nakatsugawa et al., 1965; Wong and Terriere, 1965). Endrin incubated with pig or rat liver microsomes in the presence of reduced nicotinamide adenine dinucleotide phosphate (NADPH) gave a monohydroxy-derivative, the formation of which was inhibited by sesoxane, indicating MFO involvement (Brooks, 1969). It soon became clear from mammalian studies that the ­ inversion of the unchlorinated norbornene nucleus in isodrin and endrin (as compared with aldrin and dieldrin) exposes this ring to enzymatic hydroxylation in vivo and greatly increases the rate of elimination of these compounds from mammalian tissues, in contrast to their behavior in insect tissues. Endrin is generally more toxic to vertebrates and less toxic to some insects than dieldrin; whereas the latter undergoes 9-hydroxylation syn to the epoxide ring and 9-HD (4, Figure 96.3) is eliminated by conjugation in mammals, endrin is both syn- (slowly) and anti(rapidly) hydroxylated; the anti-derivative (5, Figure 96.3) is rapidly conjugated and excreted but the syn-isomer (6, Figure 96.3) is further oxidized to 9-keto-endrin (9-KEN, also called 12-keto-endrin; 7, Figure 96.3), a remarkable example of the profound influence of stereochemistry on metabolic pathways. Bridge-end (tertiary) hydroxylation also occurs and endrin trans-dial is a minor metabolite. 9-KEN is some fivefold more toxic than endrin to rats and appears to be the ultimate toxic metabolite of endrin (Bedford et al., 1975a; Hutson et al., 1975). Species differences are evident, since Kadous and Matsumura (1982) reported the order of endrin metabolite toxicity to male German cockroaches as 5-OH anti-9-OH 9-keto-, whereas the order on topical application was 9-keto  syn-9-OH  endrin » anti-9-OH to houseflies and 9-syn-OH  9-keto-  endrin » anti-9-OH to blowflies (Brooks and Mace, 1987). Also in this report, syn-9-hydroxydieldrin (9-HD; 4) was essentially nontoxic to houseflies and blowflies, whereas the order 9-oxadieldrin (9-OD; 13, Figure 96.1)  dieldrin  9-ketodieldrin (9-KD; 8, Figure 96.3) and 9-oxadieldrin (9-OD)  9-KD  dieldrin, respectively, was found for houseflies and blowflies. Toxic 9-KD is apparently not formed from 9-HD in vivo, possibly because, in contrast to the situation with endrin, steric hindrance prevents enzymic attack on the hydroxyl group. Each set of toxicities lies within a narrow range and the toxicities of 9-KD and 9-OD might be expected to be similar, because 9-OD is an isostere of 9KD, in which C5O has been replaced by the more compact 5,8-bridged oxirane structure. These results show that several of the oxidative metabolites of these insecticides

Chapter | 96  Interactions with the Gamma-Aminobutyric Acid A-Receptor

retain insect toxicity and may contribute to the toxic effect of the parent insecticides.

2071

not formed in vivo but can be obtained indirectly by chemical synthesis. The biotransformations of chlordene and heptachlor involve allylic hydroxylation for chlordene, hydrolysis of allylic chlorine for heptachlor, epoxidation (Miles et al., 1969), and epoxide ring hydration (Brooks, 1966; Figure 96.4). Microorganisms can degrade heptachlor by removing the allylic chlorine, either reductively or by hydrolysis, so that the degradation routes for chlordene can then be followed (Miles et al., 1969); in some soils, the production of 1-hydroxychlordene (1, Figure 96.4) is comparable to HE160 production. The hydroxylated metabol­ ites appear to be detoxification products in mammals. This is difficult to prove in insects, however. Chlordene (2, Figure 96.4) and its exo-epoxide (3, Figure 96.4) have a weak housefly toxicity, which is synergized 10-fold by sesoxane, suggesting that the biotransformations observed in microsomal preparations are detoxications (Brooks, 1966; Brooks and Harrison, 1964a, 1967a,b). Is heptachlor much more toxic than chlordene because the allylic chlorine inhibits hydroxylation in this position and also ensures that heptachlor is converted into the metabolically stable epoxide? This is a question reminiscent of the aldrin/dieldrin situation. The view that heptachlor is intrinsically toxic is supported by the toxi­city (Table 96.3) of the alpha- and beta-dihydroheptachlor isomers (Figure 96.4), formed by the addition of hydrogen

96.2.3  Heptachlor, Chlordene, Dihydroheptachlor, Chlordane,   and Isobenzan Further chlorination of the feebly toxic chlordene, the Diels-Alder adduct of “hex” and cyclopentadiene, gave heptachlor, the dihydroheptachlor isomers (Table 96.3 and Figure 96.4), and the chlordane isomers (Figure 96.5). The nontoxic adduct of “hex” and cis-2-butene-1,4-diol, namely 5,6-bis(hydroxymethyl)-hexachloronorbornene-2-ene, is the precursor to which isobenzan, endosulfan (Figure 96.6), bromocyclen (Bromodan), and chlorbicyclen (Alodan) (Figure 96.7; 20 and 21, respectively) are related. The last two compounds were once used to control animal ectoparasites because of their low mammalian toxicity; endosulfan is still used extensively today, whereas isobenzan was discontinued in 1965. A preparation of heptachlor is mentioned in the original Hyman patent (Hyman, 1949) on chlordane. Numerous investigations from 1951 demonstrated the formation of heptachlor exo-epoxide, m.p. 160°C (HE160; Figure 96.4); the less insecticidal endo-epoxide, m.p. 90°C (HE190) is

Table 96.3  Toxicities of Dihydroheptachlor and Chlordane Isomers to Housefly and Mouse A, exo– B, exo–

D 6C1 H

C, endo–

Compound

A

B

C

D

Housefly LD50 (g/fly)a

Housefly LD50 (g/fly with sesoxane)a

Mouse acute oral LD50 (mg/kg)

(1) -

Cl

H

H

H

0.26

0.015

1,285

(2)  -

H

Cl

H

H

0.16

0.015

9,000

(3)  -

H

H

Cl

H

1.8

0.07

6,000

b

Cl

H

Cl

H

0.22

0.22

1,100

b

(5)

Cl

Cl

H

H

0.08

0.08

500–600

(6)

H

Cl

Cl

H

0.04

—

600

Cl

H

H

Cl

0.04

—

31

(4)

b

(7)

Alodan

—

0.31

0.05

15,000c

Dieldrin

—

0.02

0.02

75–100

Data compiled from Brooks and Harrison (1964a, 1967b) and Buchel et al. (1966a,b). a Topical application; sesoxane applied (5 g/20 mg fly) before insecticide. b 4, trans-chlordane; 5, cis-chlordane; 7, -chlordane. c LD50 for rat.

Hayes’ Handbook of Pesticide Toxicology

2072

OH

O

OH

O

PI (I) RA (I)

M

O

CI

O PI (I) RA (I)

OH

OH

H(v) M H(v) (I) (I) B PI (I) OH

CI OH

CI

OH

H(v) (I)

(3)

PI (I) RA (I)

OH

O

PI (i)

HE 190

PI(I) RA(I)

R(v)

(1)

PI(I) RA(I)

R(?) H(v) (I)

E 6CI CI

CI

CHLORDENE (2) HEPTACHLOR SYNTHESIS (HCI GAS)

CI 6

H(I)

I (v) (I); M(v) (I); M(V) CI

ALPHA-

R(V) PI (I) R(V) OH

DIHYDROHEPTACHLOR ISOMERS

+

CI +

BETA-

O HE 160

PI(I)

GAMMA-

CI

PI(I) H(I) CHLOROHYDRINS, ETC. AS FOR BETA-ISOMER

OH

CI OH

O

Figure 96.4  Biotransformations of chlordene, heptachlor, and the dihydroheptachlor isomers. B, bacteria; H, housefly; I, some insects; M, mammals generally; Pi, pig; R, rat; m, microorganisms; Ra, rabbit; E, abiotic conversion; v, in vivo; i, in vitro. All structures contain the fully chlorinated norbornene moiety.

chloride to the double bond of chlordene. Their housefly toxicity also is synergized by sesoxane, which suggests that metabolic hydroxylation, which for them replaces the epoxidation of heptachlor, results in detoxication. Beta-dihydroheptachlor (beta-DH; Figure 96.4; 2, Table 96.3) is particularly interesting because of its low mammalian toxicity (Buchel et al., 1966a,b). In the presence of NADPH, pig liver microsomes converted alpha-, beta-, and gamma-DH into a variety of hydroxylation products, which are illustrated for beta-DH in Figure 96.4. These were chlorohydrins, obtained by simple hydroxylation of the cyclopentane rings; alcohols, formed by elimination of the single chlorine atom on the cyclopentane ring; dihydroxycompounds; and a ketone (e.g., 2-keto-dihydrochlordene from beta-DH, which may afford the corresponding alcohol via a keto-reductase reaction). The 2-OH-dihydrochlordene excreted by rats fed beta-DH (Korte, 1967) may arise in this way. Similar metabolites were produced in housefly

microsomes, although no dihydroxy-compounds were detec­ted. Sesoxane inhibited the hydroxylations, which doubtless explains the synergism against houseflies observed in vivo (Table 96.3). The metabolism of the two chlordane isomers, alpha( trans-1,2-dichlorodihydrochlordene) and beta- ( cis1,2-dichlorodihydrochlordene), is complex (Figure 96.5). Either isomer might give heptachlor by dehydrochlorination and hence HE160 and all the metabolites arising therefrom. In fact, the metabolites in rats include 1-exo,2dichlorochlordene (1, Figure 96.5), oxychlordane (2), 1-exo-hydroxy-2-chlorochlordene (3), 1-exo-hydroxy2-chloro-2,3-epoxychlordene (4), 1-exo-hydroxy, 2endo-chlorodihydrochlordene (chlordene chlorohydrin), 1,2-trans-dihydroxydihydrochlordene, and the metabolites of heptachlor (Brimfield and Street, 1979; Brimfield et al., 1978; Tashiro and Matsumura, 1977). A similar series of compounds was excreted in the form of unidentified

Chapter | 96  Interactions with the Gamma-Aminobutyric Acid A-Receptor

OH (4)

O CI HEPTACHLOR, 3-OH-CHLORDANE

R(v) OH (3)

R(I) RA(v)

CI

R A(v) R(v) CI6

CHLOROHYDRINS

RA(V)

R(V) CI6

CI 1

3

CI

ALPHA-(TRANS-) 2 CHLORDANE

BETA-(CIS-) CHLORDANE

R(v) RA(v)

P(I)

2,3-EPOXIDE ‘OXYCHLORDANE’ R(v) (2)

1 CI 3

2

CI

R(v) CI (1) CI

Figure 96.5  Biotransformations of the cis- and trans-chlordane isomers. Abbreviations as in Figure 96.3. All structures contain the fully chlorinated norbornene moiety.

conjugates in the urine of rabbits treated with these chlordane isomers (Balba and Saha, 1978). These biotransformations demonstrate the remarkable versatility of the drug-metabolizing enzymes. In particular, the formation in rats of oxychlordane (Figure 96.5), analogous to heptachlor epoxide and said to be more toxic than trans-chlordane (Street and Blau, 1972), is a bioactivation due to the unexpected formation of a stable epoxide in vivo, presumably following an enzymatic desaturation that introduced a 2,3-double bond. There was no evidence for epoxide formation from either cis- or trans-chlordane in houseflies, however. Notably, trans-chlordane was threefold less toxic than cis-chlordane to this insect, and neither isomer was synergized by sesoxane (Brooks and Harrison, 1964a), indicating that intrinsic toxicities were being measured. Alpha- and beta-DH were as toxic as heptachlor when synergized; synergized gamma-DH was fourfold less toxic than synergized alpha- and beta-DH and as toxic as cis-chlordane (Table 96.3). This suggests that the 2-endo-chlorine atoms in trans-chlordane and gamma-DH contribute less to toxicity than the exo-chlorines present in alpha- and beta-DH and cis-chlordane. Moreover, an additional chlorine introduced into the 2-exo-position of gamma-DH (Table 96.3) increases its toxicity more than 40-fold, so that the resulting gem-dichloro-compound is as toxic as beta-DH having the single exo-chlorine in this position. Does this extra exo-chlorine simply reduce the possibilities for metabolic detoxification that are more likely for gamma-DH (exo-side of the ring exposed to enzymatic

2073

attack) and trans-chlordane, or do exo-chlorines increase the affinity of these molecules for a critical binding site in the nervous system? That synergized gamma-DH is as toxic as cis-chlordane (unaffected by sesoxane) may suggest that metabolism is the only factor involved and that the exo- or endo-disposition of the chlorines is immaterial. There is also the interesting question of the role of symmetry; the most insecticidal compounds in this series are beta-DH, the 2,2-gem-dichloro-analog (gamma-DH), and 1-exo, 3-exo-dichlorodihydrochlordene (delta-chlordane; Table 96.4), all having a plane of symmetry, in contrast to the other molecules discussed, for which the enantiomers may differ in toxicity (see later discussion on the heptachlor epoxide enantiomers in Section 96.3.1). Production of isobenzan (Telodrin) ceased in 1965 (Jager, 1970), but this molecule (Figure 96.6) remains of theoretical interest as acyclic ether analogue of gammachlordane, which, like the latter, has high insect and mammalian toxicity. Enzymatic attack on the chlorinated cyclic ether structure of isobenzan analogous to the biotransformations noted for the chlordane isomers results in hydrophilic metabolites such as derivatives of the gammahydroxy-acid (1, Figure 96.6), which afforded the lactone (2) and alcohol (3) on hydrolysis. Alternatively, these might arise directly by oxidative or hydrolytic elimination of chlorine atoms from the cyclopentane ring. Of particular interest because they illustrate the variety of structures having toxicity in this series is the mixture of two interconvertible isomeric ketones (14 and 15, Figure 96.7), with high insect and mammalian toxicity (housefly LD50, 0.5 g/g; rat LD50, 7 mg/kg), which can be obtained chemically from 5,6-bis(hydroxymethyl)-HCNB (“endosulfandiol”). Transannular dehydrochlorination affords an even more toxic cage ketone (16, Figure 96.7; housefly LD50, 0.25 g/g; rat acute oral LD50, 1.0 mg/kg). These analogues of isobenzan are more compact versions of the various cage molecules formed from dieldrin and provide further evidence that the dichloroethylene moiety of cyclodienes can be replaced by other polar moieties without loss of toxicity and with increased toxicity in some cases.

96.2.4  Endosulfan (Thiodan) Technical endosulfan is a 7:3 mixture of the alpha- (m.p. 109°C) and beta- (m.p. 213°C) isomers, the former (Figure 96.6) having an “extended,” dieldrin-like structure (see also Section 96.4.3) and the latter having a cage-like structure resembling endrin stereochemically. The alpha-isomer is more toxic than the beta-isomer to mammals and houseflies; both are oxidized in vivo to endosulfan sulfate (4, Figure 96.6), which resembles beta-endosulfan stereochemically and has similar toxicity to alpha-endosulfan, so that this conversion is analogous to aldrin-to-dieldrin conversion. The cyclic sulfite (and sulfate) ester structures

Hayes’ Handbook of Pesticide Toxicology

2074

Table 96.4  Insect Toxicity of Aldrin and Dieldrin Relatives, Including Some Molecules with Fewer Chlorine Atoms a S

1O

1

8a

2

Y6 X = Y = carbon, except for compounds (3) and (4)

9 S

a

3

Chemical

5

4a

4

X7

8

Chlorination in aldrin analog

HFa

GRb

1

2

3

4

10-syn

10-anti

(1)

---------

---

---

6-Cl

----- -

------(aldrin)

0.55

2.0

(2)

---------

---

---

6-Cl

----------- -

6,7-epoxide: dieldrin

1.0

1.0

(3)

---------

---

---

6-Cl

----------- -

6,7-N  N-

3.6

5.3

(4)

---------

---

---

6-Cl

----------- -

6,7-NN(→O)–

4.45

2.1

(5)

1

H

H

4

10s

10a

2.1

9.3

(6)

1

H

H

4

10s

10a (6,7-epoxide; DD)

3.8

8.0

(7)

1

H

H

4

H

H

0.03

0.8

(8)

1

2

3

4

H

10a

0.08

0.65

(9)

1

2

3

4

10s

H

0.02

Inactive

Data compiled from Soloway (1965). a Housefly toxicity compared with dieldrin (1.0) by direct spray. b German cockroach toxicity compared with dieldrin (1.0) by exposure to dry films on paper.

6CI

OH mfo ? O

O

O

O

SO2

(3)

ENDOSULFAN ‘ETHER’

(4) ENDOSULFAN SULFATE HYDROLYSIS

OXIDOREDUCTASE ?

LOCUST, MOUSE, HOUSEFLY

mfo

CI

CI CH2OH CH2OH ENDOSULFAN-DIOL

ENVIRONMENT BIOTIC, ABIOTIC

CI

O O

CI CI

CI

O

6CI

LOCUST S

O

O O

(2) ENDOSULFAN ‘LACTONE’

ALPHA-ENDOSULFAN

MOUSE, RAT

CONJUGATION

6CI CI CH2OSO3H COOH (1)

HYDROLYSIS

CI

O

ISOBENZAN

Figure 96.6  Major transformations of alpha-endosulfan and isobenzan. All structures contain the fully chlorinated norbornene moiety. Note that--O-indicates the skewed (“trans”) position of the second oxygen in the “twist-chair” (asymmetric) configuration of alpha-endosulfan (Schmidt et al., 1997).

Chapter | 96  Interactions with the Gamma-Aminobutyric Acid A-Receptor

CI

CI

CI

CI

CI2HC

O

CH2CI

CI CI CI

CI CH2CI

(16)

CH2CI

CIH2C

CI CI

CI

O

CI CI H CH2CI CI (17) Toxaphene

CI

CI

6

CI

CH2BR

CH2CI (19)

(18) Toxaphene

CI

H CI

CI

(14,15)

CHCI2

CI

CI

O

H CI

O

CIH2C

CI

CI

2075

CI 6

(20) Bromodan®

CI

12

10

CH2CI CH2CI Alodan®

(21)

(22) Mirex

CI H

(23) O Chlordecone (Kepone)

CI O

CI H H

(24)

Figure 96.7  Chemical structures of compounds mentioned in the text.

completely alter the behavior of the endosulfans, which disappear quite rapidly from living tissue, partly by hydrolysis to the parent nontoxic endosulfan-diol and metabolites similar to those formed from isobenzan. The sulfate is formed faster from the alpha- than from the beta-isomer in houseflies and is as toxic as beta-endosulfan to these insects (Barnes and Ware, 1965); cyclodiene-resistant flies eliminated these isomers more rapidly than normal (S-) flies, but the tissues contained only the toxic sulfate, which also appears in the body fat of mammals but disappears rapidly when exposure ceases. Endosulfan-treated locusts excreted the sulfate, endosulfan ether, alpha-hydroxy-endosulfan ether (3, Figure 96.6), and the corresponding lactone (2). Endosulfan-treated mice stored the sulfate transiently in their fat and excreted endosulfan, the sulfate, and the parent diol in feces (Maier-Bode, 1968). It is evident that endosulfan is a relatively nonpersistent compound in mammals (Dorough et al., 1978) and has generally favorable environmental properties, apart from high fish toxicity,

which requires caution in aquatic situations. With the exception of the toxic sulfate, metabolites of endosulfan isomers are undoubtedly detoxication products.

96.2.5  Toxaphene, Mirex, Chlordecone (Kepone) Toxaphene (camphechlor) is a complex mixture of some 177 compounds obtained by chlorinating camphene to a 67–69% chlorine content (Pollock and Kilgore, 1980; Saleh et al., 1979). The identified compounds are actually chlorinated bornanes arising from the Wagner-Meerwein rearrangement of the camphene skeleton, among which the octachloronorbornanes; 2,2,5-endo,6-exo–8,8,9,10octachloro-norbornane (17, Figure 96.7) and 2,2,5-endo,6exo-8,9,9,10-octa-chloronorbornane (18, Figure 96.7), are highly potent, with mouse ip LD50 values of 2–3 mg/ kg (Turner et al., 1977). The less toxic 2,2,5-endo,6exo,8,9,10-heptachloronorbornane (19; compound B; LD50,

2076

75 mg/kg) was potentiated eightfold by PBO administered prior to the insecticide, suggesting the possibility of oxidative detoxication mechanisms for this compound. Experiments with rat liver preparations confirmed metabol­ ism by MFO, and the formation of glutathione and glu­ curonide conjugates (Chandurkar and Matsumura, 1979) could be demonstrated (see Smith, 1991). The positioning of the added chlorine substituents in compound B seems to be critical; at the 3-exo-position and in the 10-chloromethyl moiety, an additional chlorine greatly reduces mouse toxicity, as does the combination of 3-exo-chlorination and 5,6-dehydrochlorination, to give a vinylic chlorine atom. Notably also, the simpler (less bulky) molecules hexachloronorbornene-2,5-diene and heptachloronorborn-2-ene used to prepare cyclodiene insecticides lack toxicity, which only appears when halomethyl groups are introduced into the nucleus as in bromocyclen (Bromodan; 20, Figure 96.7) and chlorobicyclen (Alodan, 21). Both are synergized 10-fold by sesoxane in houseflies (Brooks and Harrison, 1964a; Table 96.2) and are quite good insecticides with very favorable mammalian toxicity (rat acute oral LD50s, 13,000–15,000 mg/kg); that is, they appear to be considerably more selective (insect versus mammal) than the most toxic components of toxaphene. Mirex (22, Figure 96.7) is the fully chlorinated cage molecule, formed by the self-condensation of two molecules of “hex,” and might be expected to be rather resistant to enzymatic attack. Animal tissue levels plateau only slowly on exposure and decrease very slowly when exposure ceases. One chlorine atom is reductively replaced in the environment to give photomirex (8-monohydro-mirex), which appears to behave like mirex in the rat (Chu et al., 1979; Hallett et al., 1978). Reductive dechlorination can occur in vivo; 2,8-dihydromirex and 5,10-dihydromirex have been identified as rat metabolites. Whereas 2,8dihydromirex does not appear to be further metabolized, 5,10-dihydromirex appears to be converted into more polar metabolites, which appear in rat urine (Yarbrough et al., 1983). Mirex has low mammalian toxicity (rat oral LD50 ranging from 600 to 3000 mg/kg) and its signs of poisoning differ from those produced by the less chlorinated cyclodienes. The metabolism of mirex in houseflies is equally slow, and Shankland (1982) compared its slow insecticidal action with the delayed onset of dieldrin poisoning discussed earlier. The onset of poisoning following topical application of lethal doses of mirex to the American cockroach occurred only after 3 days. Moreover, when isolated sixth abdominal ganglia were irrigated with suspensions of 5  104 M mirex for 4 h, there was no change in the patterns of spontaneous activity or elicited postganglionic responses. Ganglia excised from symptomatic cockroaches showed, however, spontaneous after-discharge behavior characteristic of poisoning following dieldrin treatment. Hemicholinium-3, which depletes Ach stars, eliminated

Hayes’ Handbook of Pesticide Toxicology

the neuroactivity in giant fibers, but the ganglia remained responsive to nicotine, as is found in dieldrin poisoning. Because mirex appears to be highly resistant to biotransformation, Shankland concluded that the delayed action was unlikely to involve a requirement for bioactivation and must arise from the intrinsic properties of this highly chlorinated molecule, such as slow penetration through diffusion barriers in the insect central nervous system. Chlordecone (Kepone, 23, Figure 96.7) differs from mirex in having a carbonyl group, which is probably responsible for its moderately rapid clearance from animal tissues. In humans and pigs, this is via the alcohol (chlordecol), and a cytosolic keto-reductase, which can effect this reduction, has been found in gerbil and human liver (Molowa et al., 1986). Bloomquist and Shankland (1983) found that chlordecone produced the same signs of poisoning as mirex in the American cockroach and concluded that chlordecone has the same mode of action as dieldrin, although, like mirex, it acts more slowly. From experiments on the displacement of [3H] picrotoxinin (PTX) binding by mirex and chlordecone from American cockroach head membranes, Tanaka et al. (1984) concluded that chlordecone interacts with the PTX-binding site, as expected, whereas mirex was much less potent in this respect; moreover, dieldrin-resistant German cockroaches were resistant to chlordecone but not to mirex. Chlordecone is also known to have inhibitory effects on neurotransmitter uptake in mammals and such an action may also contribute to its insect toxicity.

96.3  Structure–toxicity relationship and mode of action 96.3.1  Fully Chlorinated Cyclodienes: Substituted Hexachloronorbornenes (HCNB) Soloway (1965) published a comprehensive review on the structure–activity relationships of cyclodiene insecticides at a time when information on their metabolism was just beginning to appear, so his review makes only passing reference to the possible influence of metabolism but provides a great deal of information about toxicity trends in numerous series of cyclodiene analogues. Initially, he emphasized the similarity between heptachlor epoxide HE160/cis-chlordane and HE90/trans-chlordane (Figures 96.4 and 96.5), each pair having two similarly oriented electronegative atoms (i.e., oxygen and chlorine), with toxicity greater in the first (exo,exo) orientation than in the second (exo,endo) orientation of these substituents. Deltachlordane (Table 96.3) with 1-exo,3-exo chlorine substituents is a highly insecticidal symmetrical variant of the orientation found in the HE160/cis-chlordane pair. Interestingly, delta-chlordane is an analogue of alodan in which the two side-chain chlorines have become fixed

Chapter | 96  Interactions with the Gamma-Aminobutyric Acid A-Receptor

in the exo-positions by the extra carbon atom of the cyclopentane ring and their insect toxicities are of the same order when alodan is synergized by sesoxane (Table 96.3). The second (3-exo) chlorine in delta-chlordane has a severe effect on mammalian toxicity, because this molecule has a much higher rodent toxicity than either alodan, alpha- (1, Table 96.4) or beta-dihydroheptachlor (2) (DH), or the chlordane isomers (4, 5). As noted already, the order of housefly toxicity of the DH-isomers is beta-DH  alphaDH  gamma-DH; beta-DH has an exo-chlorine and is also symmetrical, alpha-DH has an exo-chlorine but is asymmetrical, whereas gamma-DH has an endo-chlorine, which, being “hidden” beneath the ring system, may be less accessible to a critical binding site and also leaves the exo-face of the ring more exposed to metabolic attack from the exo-side (compare the metabolism of endrin in Section 96.2.2). Soloway presented insect toxicity data for many derivatives of HCNB of the chlordane, isobenzan, endosulfan, aldrin, and isodrin series, together with lindane, which already appeared to have the same mode of action (Busvine, 1964). He concluded that high insecticidal activity required the presence of two electronegative centers within a narrow range of distance and direction with respect to one another and placed on or across the plane of symmetry defined by the CCl2-bridge. Many cyclodienes fulfill these requirements but some, such as dihydroaldrin, dihydroisodrin, and photoisodrin (9, 10, and 11, respectively. Figure 96.1) (only one electronegative center) and bromodan (20, Figure 96.7), alpha-DH, exo-chlordene epoxide (Figure 96.4), and HCE (8, Figure 96.1) (asymmetrical), do not, yet are indicated to have high intrinsic toxicities when their metabolism is suppressed in vivo. Evidently, the involvement of a second electronegative center such as an epoxide ring in binding to the site of action may increase the affinity of the molecule for this site, by hydrogen bonding, for example. Thus, lack of a second electronegative center may explain the earlier noted slow action of the dihydro-compounds, which, in the absence of an inhibitor of metabolic oxidations, may afford them increased opportunity for both detoxication and binding to inert storage sites. Notably, the cage molecule photoisodrin is more polar than the related dihydroisodrin and dihydroaldrin and acts rapidly, especially when synergized; it may have the more favorable pharmacokinetic properties of the more rapidly acting epoxides, although apparently lacking their second electronegative center (Brooks, 1973; Brooks and Harrison, 1963). There is limited information about the relative toxicities of enantiomeric forms of chiral cyclodienes, which are obviously of interest in this context. The epoxide hydrolases of pig liver microsomes selectively hydrate the same enantiomers of chlordene epoxide, HCE and HE90 (Brooks et al., 1968). The isolated residual epoxides appeared to have the same order of toxicity to houseflies as their

2077

respective racemates, which are not detectably hydrated by this insect (Brooks et al., 1970). Miyazaki et al. (1978, 1979, 1980) synthesized the pure enantiomers of chlordene, chlordene exo-epoxide, HE160,2-chloroheptachlor (Figure 96.8), and 3-chloroheptachlor and found that their toxicities to the German cockroach (topical LD50, g/ g) were in the order ()-chlordene (148)  racemic chlordene (300)  ()-chlordene (inactive); ()-chlordene epoxide (74)  racemic chlordene epoxide (158)  ()epoxide (inactive); racemic heptachlor (2.64)  ()-heptachlor (3.38)  ()-heptachlor (5.32); ()-HE160 (1.29)  racemate (1.82)  ()-HE160 (2.98); ()-2-chloroheptachlor (20)  racemate (50)  ()-2-chloroheptachlor (1 0 0); 3-chloroheptachlor (enantiomers and racemate inactive). Miyazaki et al. concluded that ()-chlordene is intrinsically nontoxic to the German cockroach, observing that the corresponding ()-epoxide (nontoxic) formed in vivo is metabolized to the expected oxidative and hydrolytic

Cl6

Cl6 O (+) –

(–) –

Cl6

Cl6

O (–) –

(+) –

Chlordene exo-epoxide

Chlordene

Cl6

Cl6

O C1

C1 (–) –

(–) –

Cl

Cl6 (+) –

Cl

Cl6

1

O

2

(+) – Heptachlor exo-epoxide (HE 160)

Heptachlor Cl6

Cl6

Cl

Cl (–) –

Cl

Cl (+) –

2-Chloroheptachlor Figure 96.8  Absolute stereochemical configuration of the enantiomers of chlordene, chlordene epoxide, heptachlor, heptachlor epoxide (HE160), and 2-chloroheptachlor as established by Miyazaki et al. (1978, 1979, 1980).

2078

Hayes’ Handbook of Pesticide Toxicology

products (Brooks and Harrison, 1965) at about the same rate as the toxic ()-epoxide from observably toxic ()chlordene. However, they also considered ()-chlordene to be intrinsically inactive, therefore requiring bioactivation by conversion into the toxic ()-epoxide in vivo. Unfortunately, these experiments did not include a synergist to suppress oxidative metabolism. Experiments with houseflies showed that both chlordene and dihydrochlordene had low but measurable toxicities to that insect, which were synergized by sesoxane (Brooks, 1966; Brooks and Harrison, 1964a); in fact, synergized chlordene was only fivefold less toxic than synergized chlordene exo-epoxide. The role of epoxidation in the toxicities of the heptachlor enantiomers (or those of 2-chloroheptachlor) has not been reported. The ()- to ()-heptachlor toxicity ratio for German cockroach was 1.56; for ()- to ()-HE160 it was 2.3 and for ()- to ()-2-chloroheptachlor it was 5.0, with the more toxic ()-antipodes and toxic ()chlordene epoxide all having the same absolute stereochemistry (Figure 96.8). Apart from ()-chlordene epoxide and the anti­podes of 3-chloroheptachlor, the other antipodes are clearly all active but the ratio of 2.3 for the HE160 antipodes is likely to be the safest measure of comparative intrinsic toxicities in this series because the known stability of this epoxide should avoid or minimize the complication of metabolism in vivo. Thus, although one absolute configuration of HE160 is favored, both are toxic, which might be expected if the critical binding site is in a symmetrical (or nearly symmetrical) cylinder of about the same diameter as the molecules discussed, so that either antipode can interact reasonably well with such a site in the bore of the structure, now known to be the chloride ionophore of the GABAA-receptor (Section 96.4). Notably, alpha-DH must exist in enantiomeric forms, which, if superimposed, give a symmetrical “composite” molecule that resembles both delta-chlordane and isobenzan (its oxygen isostere). Likewise, superimposition of the enantiomers of both HE160 and HE90 gives “composites” that are similar to both delta-chlordane and isobenzan. Such symmetrical molecules might be expected to interact particularly well with a close-fitting cylindrical binding site.

the chlorine atoms in the hexachloronorbornene moiety could be replaced by hydrogen (Table 96.4). Species differences were evident; an aldrin analogue (7, Table 96.4) having only the two one- and four-bridge chlorines was reported to be nearly as toxic as dieldrin to the German cockroach, although nontoxic to other insects tested. In aldrin, the methano-bridge chlorine atom anti to the chlorinated double bond was found to be more important for toxicity than the syn-chlorine (compare 8 and 9, Table 96.4), and replacement of the two ethylenic chlorines in dieldrin by hydrogen to give didechloro-dieldrin (DD; 6, Table 96.4) increased housefly toxicity fourfold and toxicity to the German cockroach eightfold; the latter insect appears to be particularly sensitive to these compounds. The high toxicity of diazaaldrin and its N-oxide (3 and 4, Table 96.4) should also be noted. Of interest was the possibility that if the increase in toxicity effected by replacement of the ethylenic chlorines in dieldrin proved to be a general phenomenon for cyclodiene insecticides, it might be possible to combine the change to a tetrachloronorbornene moiety with a more labile epoxide ring or other labile system (e.g., cyclic sulfite as in endosulfan) to produce useful insecticides having both oxidative and hydrolytic detoxication routes that would be more selective and environmentally acceptable. Selective dechlorination of several series of cyclodienes was undertaken to test this possibility (Brooks, 1975, 1977, 1980, 1985; Brooks and Mace, 1987; Brooks et al., 1981). It transpired that the effect of dechlorination was not uniform but depended on the structure of the molecule as a whole. The most consistent observation for all series was the greater importance of the anti- versus the syn-chlorine atom in the pentachloronorbornene moiety, as noted for aldrin by Soloway (1965). These two changes for dieldrin were combined to give the 1,4, anti-10-trichloro-analogue of dieldrin (DSD; 24, Figure 96.7), which approached dieldrin in toxicity to houseflies and blowflies and was sevento 20-fold more toxic than its 10-syn-chloro-isomer with the chlorine atom adjacent to the double bond. Thus, the pentagonal arrangement of chlorine atoms evident in lindane and in the cyclodienes derived from HCNB (Busvine, 1964) is not sacrosanct for cyclodienes.

96.3.2  Compounds with Fewer,   or No Chlorine Atoms

96.3.2.2  Structural Convergence of Cyclodienes and their Dechlorinated Analogue with Other Cage Convulsants Acting at the Chloride Ionophore

96.3.2.1  Reductive Dechlorination of Cyclodienes Early information (Soloway, 1965) indicated that the unchlorinated methano-bridge of aldrin could be replaced by 9-syn-Cl-CH- and that of isodrin by -CH2CH2- or spiro­ cyclopropane, but the overall molecular length could not exceed that delineated by dieldrin or alpha-endosulfan. There were, however, interesting indications that some of

Further information arose from comparisons with the naturally occurring convulsant picrotoxinin (PTX; Figure 96.9), which Hathway et al. (1965) found to have effects similar to those of dieldrin and isobenzan on ammonia metabol­ ism in rat brain. PTX antagonizes the action of GABA by blocking the chloride ion channel associated with its receptor (Kadous et al., 1983; Takeuchi and Takeuchi,

Chapter | 96  Interactions with the Gamma-Aminobutyric Acid A-Receptor

O

CH3

O

A (trans–) CH3 CH2 H H β S (cis–) OH

γ α

H

O

O

O

γ

H

e a e H γ β e H H aH a C1 α e e C1 H C1 C1 H C1 a a e

β

α

C1 C1

O

a C1

A C1 S C1

C1

H

2079

C1 C1

C Lindane

B Cyclodienes

A PTX

anti-10-Cl

trans-isopropenyl

A

Cl1 α

β S

Cl4 γ

s H H

α

A β γ

H

O O C12

C13

D

E

Figure 96.9  Structures of (A) picrotoxinin (PTX) with its bulky beta-isopropenyl group trans- (or anti-) to the lactone ring; (B) heptachlor epoxide; (C) lindane, showing the aaaeee configuration of chlorine atoms essential for toxicity. A, anti-; S, syn-; a, axial; e, equatorial substituents. D shows the highly chlorinated face of a cyclodiene insecticide, as in heptachlor epoxide (B above), and E, the lactone ring system as in PTX, both viewed from the right (“end-on” position). In the corresponding view of lindane (C viewed from below), note that the chlorine equivalent to the syn-10-chlorine of fully chlorinated cyclodienes is replaced by a hydrogen (double arrow). However, the electrostatically more favored superimposition of lindane on PTX (Calder et al., 1993) is C viewed from the right, in which the three-axial chlorines of lindane are together equivalent to the trans-bulky isopropenyl substituent and lactone ring of PTX.

1966, 1972). Evidence was then reported that cyclodienes and lindane compete with PTX at a common binding site in cockroach brain (Matsumura and Ghiasuddin, 1983; Tanaka et al., 1984), and this was the site of convulsant action of these compounds, a proposal supported by the similarity in the neurophysiological effects of the cyclodienes and PTX and the cross-resistance to PTX shown by cyclodiene-resistant cockroaches. The structural similarities among PTX, HE160, and lindane led Ozoe and Matsumura (1986) to elaborate the two electronegative center hypothesis of Soloway (1965) in a series of PTX analogues that emphasized the importance of the bulky trans-substituent on the lactone ring as a third requirement for interaction with the PTX binding site. This is the point of convergence (Figure 96.9) with the cyclodienes (B) and lindane (C), for which the 10anti-chlorine and the central axial chlorine (or the central equatorial chlorine), respectively, appear to provide the appropriate bulky substituent (Brooks and Mace, 1987) and the highly toxic cage compounds such as t-butylbicyclophosphorothionate (TBPS) and t-butylbicycloorthobenzoate (TBOB), in which the t-butyl-moiety is the necessary bulky substituent (Casida et al., 1985; Palmer and Casida, 1985). The highly insecticidal orthobenzoate EBOB (Figure 96.10) proved to be the best ligand for the insect GABA-receptor chloride ionophore, and [3H]-EBOB

has been used subsequently in binding displacement studies with numerous putative channel blockers at the PTX binding site. Attempts to simplify the orthobenzoate ring system of these potent cage convulsants and the search for structural changes to confer selective toxicity in favor of mammals versus insects produced numerous insecticidal dithianes, oxathianes, and their sulfoxides and sulfones (Palmer and Casida, 1995; Wacher et al., 1992). Figure 96.10 further shows the convergence of structural changes between partly (Brooks and Mace, 1987) and totally (Ozoe et al., 1990) dechlorinated alpha-endosulfan (1, Figure 96.10) (which surprisingly retains measurable housefly toxicity (LD50, ca. 43  g/g; 20-fold less toxic than alpha-endosulfan, when co-applied with sesoxane), the toxic t-butyl trioxabicyclooctanes (TBOs) such as 2 (Figure 96.10) (Palmer and Casida, 1985), and the dithianes described by Elliott et al. (1992). Although the 1,3-dioxan (not shown but analogous to structure 7) corresponding to 2 (Figure 96.10) is weakly insecticidal (Palmer and Casida, 1995), the hybrid molecule (3) of deschloroalpha-endosulfan (1) and 1,3-dioxan was more toxic than fully dechlorinated alpha-endosulfan (1), indicating that extra rigidity conferred on the 1,3-dioxan structure can improve insecticidal activity. In the hybrid molecule (3), the norbornene moiety appears to provide the bulky substituent (compare t-butyl in 2), while possibly increasing

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Cl

anti– 4Cl

Cl

O O

H syn–

S

4H

O O

H

0

anti - 10, pentachloroisomer (1.0)

O 6H

O O

P=S TBPS

O S O 0 (>500)

Cl 4H

O O

S

O

0

R

6H

(1) deschloro - endosulfan (43)

R= (3) 4–Br–Ph (15) (4) 4–CH ≡ C–Ph (0.33) (5) 4–CN–Ph (5.5)

O O

O

S S

6C1

R

R= phenyl (Ph) (TBOB) (2) R= 4–Br-Ph (0.8) 4–CH≡C–Ph various TBOs 4–CH–Ph O 4 1 Ph-4-C≡CH O O EBOB H axial

H

O

(6) (5.5)

O

0

(265) Cl

various dechloro endosulfans

O

S

Ph-4-C≡CH

O O (7) (0.8)

2

Ph-4-C≡CH

(8) (0.24)

Figure 96.10  Convergence between exploration of reductively dechlorinated alpha-endosulfan analogue (norbornene type) (Brooks and Mace, 1987; Ozoe et al., 1990, 1993), the trioxabicyclooctane-derived cage convulsants, and the more recent dioxans and dithianes (Pulman et al., 1996). The bracketed number following a chemical number or structure is the housefly topical LD50 (g/g; measured in the presence of sesoxane or piperonyl butoxide).

the conformational rigidity of the dioxan moiety. Further­ more, Ozoe et al. (1993) showed that the extended structure 4 (Figure 96.10) is highly insecticidal and that compound 5 has intermediate toxicity, which is entirely lost by any chlorination in the norbornene nucleus. Thus, the dioxan (5) loses its housefly toxicity (LD50, 5.5 g/g) completely when one- or two-bridge chlorines are introduced and there is evidently a crossover point between the two structural types, because the fully chlorinated but not extended dioxan (6) has the same toxicity, which is much reduced in various partially dechlorinated analogues (see Section 96.4.3 for further discussion). Bridge bis-chlorination of fully dechlorinated alphaendosulfan (Figure 96.10) greatly reduces its housefly toxicity (Ozoe et al., 1993), but the 10-anti-chlorine analogue retains measurable toxicity and several dechlorinated analogues having this Cl atom in combination with one ethylenic chlorine and the two bridge-end chlorines are highly toxic, showing that the additional chlorines are required for binding this shortened molecule in the critical site (Brooks and Mace, 1987), implying the requirement for a minimum of four chlorines for high toxicity. Nevertheless, at least one of the bridge-end chlorines in cyclodienes can be replaced, because a dieldrin analogue (25, Figure 96.11), which has one bridge-end carbon atom replaced by nitrogen, has appreciable insect toxicity (Gladstone and Wong, 1977).

Many trioxabicyclooctanes are highly toxic to both mammals and insects but remarkable selectivity can be conferred on some structures by appropriate derivatization; thus, the trimethylsilyl-derivative (26, Figure 96.11) of the 4-n-butyl analogue of EBOB (Figure 96.10) is highly toxic to houseflies (LD50, 0.43 g/g) but poorly toxic to mice (LD50,  400 mg/kg). This derivative appears to be oxidatively reconverted into its toxic ethynyl precursor in houseflies, whereas in the mouse this oxidative bioactivation is much less important (Palmer et al., 1991b). Because all the compounds mentioned and some other classes of cage convulsants are believed to act at the PTX binding site, a considerable array of compact molecules is now available with which to delineate this site. A 4ethynylphenyl-substituent in the 2-position of 5-t-butyl-1, 3-dioxane (7, Figure 96.10) or 1,3-dithiane (8, Figure 96.10) was found to be more effective than a 4-bromophenylsubstituent; the trans-(linear) ethynylphenyl dithiane (8) was somewhat more toxic to houseflies than the analogous trans-dioxane (7) and cis- (angular) isomers were generally equally toxic to or less toxic than trans-isomers (Palmer and Casida, 1995; Pulman et al., 1996). With the possibility of oxidation at sulfur in vivo, which may enforce additional conformational rigidity and also increase binding propensity, the situation becomes more complex (see Section 96.4.3).

Chapter | 96  Interactions with the Gamma-Aminobutyric Acid A-Receptor

Cl

Cl

Cl

O O

O N

PH–4–C≡CSi (CH3)3

O Cl (25)

(26)

Cl CF3 Cl

N

Cl

(30)

N

N

PH–4–C≡CH

N

(27) NC

Cl

(28)

NH2

N

Fipromil CF3

Cl

CF3

Cl

N

Si (CH3)3

(29)

SOCF3

(31)

N

PH–4–C≡CH

S

N

N CH3

N

Cl

CF3

Cl

(32) N

Cl

N

O

N SCH2F

CF3

CN NH O S O

(33) H3C

(34)

CN CF3 CF3

O N

O O

Si–PH–4–C≡CH

(35)

Cl

Figure 96.11  Chemical structures of compounds mentioned in the text (25–35).

96.3.3  Links between Polychlorocycloalkane and Recent Heterocyclics Apparently Acting at the Chloride Ionophore Recently, arylpyrazoles, such as fipronil (27, Figure 96.11), and various 5-alkyl-2-arylpyrimidines (28) and 1,3thiazines (29) (Pulman et al., 1996), in which the planar heterocyclic ring replaces the spacers formed by the TBO and 1,3-dioxane and dithiane structures, have been added to the list of chloride ionophore blockers. Insecticidal activity was also found in triazoles (30, 31) (Boddy et al., 1996; Von Keyserlingk and Willis, 1992) and pyrimidinones (32) (Whittle et al., 1995) and a spirosultam (33) (Bloomquist et al., 1993), demonstrating the diversity of structures that probably act at this site. Cole et al. (1994) examined the inhibition of [3H]alpha-endosulfan binding in housefly head membranes by lindane and several cyclodienes and concluded that these insecticides are the only GABA-receptor ionophore blockers that consistently inhibit the binding in these membranes, not only of the earlier used ligands such as [35S]-TBPS and [3H]-EBOB, but also of [3H]-alphaendosulfan. However, a representative dithiane, EBOB, fipronil, and other pyrazoles were less effective in inhibiting [3H]-alpha-endosulfan binding than the chlorinated

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insecticides, from which it appeared that the latter compete directly for the endosulfan site, whereas the others bind with different inhibition kinetics or at a site more closely coupled to the EBOB than to the endosulfan binding domain. Notably, the channel activator avermectin Ba did not inhibit endosulfan binding. An even more suitable ligand for the chlorinated insecticides is [3H]-BIDN (34, Figure 96.12) (Holyoke et al., 1994), a simple norbornene derivative, which has high insect and mammalian toxicity (Kölbl et al., 1981; Middleton and Bingham, 1982). Several putative affinity probes for the binding site have also been described. When aryl pyrazoles synthesized as herbicides were found to be insecticidal, their convulsive activity was not immediately recognized to result from GABA-antagonism (Klis et al., 1991). Cole et al. (1993) reported subsequently that several compounds, including fipronil (27, Figure 96.11) (Colliot et al., 1992; Hatton et al., 1988), which has become a commercially successful insecticide, blocked the GABA-gated chloride ionophore with higher potency for a site in housefly than in mouse brain, offering the possibility of selective toxicity. Fipronil has relatively low acute mammalian toxicity (Section 96.4.3). It inhibits [3H]-EBOB binding to housefly head membranes and dieldrin-resistant flies show some resistance to it (Cole et al., 1993; Colliot et al., 1992), providing a clue to its mode of action. The cyclodiene insecticides and lindane were found to be potent displacers of [35S]-TBPS binding to GABAreceptors in rat brain and inhibitors of GABA-dependent 36 Cl ion flux into rat brain microsacs, from which it was suggested that these PCCAs act as noncompetitive blockers of GABAA-receptors (Abalis et al., 1985; Gant et al., 1987; Lawrence and Casida, 1984). Potency in these assays correlates with toxicity (Casida et al., 1988), but TBPS is not a potent insecticide and [35S]-TBPS is unsuitable as a radioligand for insect studies; it appears that the structural features required for binding at the housefly GABA-receptor are different from those for the mammalian one and [3H]-EBOB, a highly potent insecticide, was ultimately designed as a superior ligand for insect binding studies (Deng et al., 1991) and generally provides a good correlation between its displacement by PCCAs and their housefly toxicities. By use of this ligand, it was concluded that PCCA, PTX, dithiane-related compounds, and phenyl­ pyrazoles all have the same mode of insecticidal action, a view supported by the up to 27-fold cross-resistance to EBOB shown by dieldrin-resistant houseflies (Cole et al., 1993). Moreover, the naturally occurring insecticide avermectin B1a and derived moxidectin (Fisher, 1997), which behave as GABA-agonists, stimulating rather than inhibiting chloride ion influx, are potent noncompetitive inhibitors of EBOB binding. This implies that avermectin action involves the chloride ionophore but that it is bound at a

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GABA-binding region

Chloride ion channel

HOOC

OUTSIDE LIPID BILAYER M4

M3

M1

M2

CYTOPLASM (INSIDE) α δ

3

4 1

γ

β

M2 M M2 2 M M2 2

binding site for noncompetitive blockers (cyclodienes, lindane, PTX, arylpyrazoles, etc.)

α

Figure 96.12  Schematic representation of the GABAA-receptor of mammalian brain, showing five trans-membrane glycoprotein subunits, each with their four trans-membrane helices (M1-M4), of which the M2 segments (shown as cylinders, and black circles in the plan view) are believed to form the pore of the integral chloride ion channel (MacDonald and Olsen, 1994). A modified subunit carrying cyclodiene resistance in Drosophila (Rdl) shows homology with the mammalian brain beta-subunit (Ffrench-Constant et al., 1991).

site different from that involving EBOB and PCCA; nor, in contrast to EBOB, is there cross-resistance to dieldrin, so that the channel modification that confers dieldrin resistance does not apparently involve the avermectin binding site (Deng et al., 1991). Based on ligand binding studies, Deng et al. (1993) proposed four partly associated sites in the housefly GABA chloride ionophore that are relevant to insecticidal action: site A, interacting with EBOB and its isosteres; B with TBPS and isosteres; C with phenylpyrazoles; and D with avermectins. Action at sites A and C gives similar signs of poisoning and cross-resistance to dieldrin; PCCA and some TBPS isosteres may act at both A and B. The avermectin site D is coupled in some way with A and C but not to the TBPS site B, which is also distinct from the phenylpyrazole site C. Thus, the redu­ced affinity for [3H]-EBOB binding observed in dieldrinresistant houseflies is due to its reduced affinity for the PCCA binding site, and the cross-resistance noted for TBOs, lindane, toxaphene, cyclodienes, dithianes, arylsilatranes (35, Figure 96.11), and PTX suggests that the structural modifications in the EBOB binding site are involved in resistance to all these insecticides (Hawkinson and Casida, 1993) but fortunately do not confer resistance to avermectins, which have very high toxicity against agricultural and household insect pests, phytophagous mites, and plant and animal nematodes.

96.4  Molecular mechanism   of action 96.4.1  Topography of The GammaAminobutyric Acid A-Receptor GABA is the principal neurotransmitter of the mammalian and insect central nervous system (CNS) and the insect neuromuscular junction. In mammals, baclofen-sensitive GABAB-receptors are coupled to calcium and potassium channels and the action of GABA is mediated by G-proteins. In contrast, GABAA-receptors, of interest here, are members of the super family of ligand-gated ion channels that contain a chloride ionophore (Schofield et al., 1987). Simplistically, an inhibitory GABA-ergic nerve terminal abutting on the presynaptic terminal of another nerve that releases a neurotransmitter (e.g., acetylcholine, ACh) releases GABA when stimulated. GABA then diffuses to the presynaptic terminal of the other nerve, where it binds to a GABAA-receptor, causing entry of chloride ions and resulting in hyperpolarization of the terminal and inhibition of release of the other neurotransmitter. Thus, postsynaptic stimulation of the other nerve by its transmitter (e.g., ACh) is reduced. This inhibitory mechanism explains the apparent cholinergic effects of dieldrin and lindane on American cockroach ganglia (Shankland and Schroeder, 1973;

Chapter | 96  Interactions with the Gamma-Aminobutyric Acid A-Receptor

Uchida et al., 1978), because disinhibition (blockade of a presynaptic chloride ionophore) of the presynaptic terminal of a cholinergic nerve should result in uninhibited ACh release and consequent hyperstimulation of the postsynaptic terminal, as is observed. The same basic mechanism for disinhibition may, of course, affect nerve terminals involving neurotransmitters other than ACh (Joy, 1982) with a variety of possible effects, depending on the species and on differing nerve architecture. The GABAA-receptor of human brain consists of four or five 50- to 60-kDa glycoprotein subunits, each of which contains four (M1–M4) hydrophobic domains (alpha-helices) that traverse the membrane (Figure 96.12) (MacDonald and Olsen, 1994; Schofield et al., 1987) and contribute to and stabilize the walls of the chloride ionophore. The five M2domains are believed to be arranged so as to form the 5.6Å-diameter lumen of the channel, with the side chains of their threonines and serines forming hydrophilic rings that contribute to the induction of ion flow.

96.4.2  Molecular Biology of Cyclodiene Resistance Recently, a cyclodiene resistance-conferring gene, Rdl, from the fruit fly, Drosophila melanogaster, has been cloned and shows homology with the mammalian brain beta-subunit (Ffrench-Constant et al., 1991). Dieldrin resistance was subsequently found to be associated with the point mutation alanine 302 to serine (Ala302-Ser) within the M2 membrane-spanning domain, near the site conferring charge selectivity in the closely related nicotinic AChreceptor (Ffrench-Constant et al., 1993a,b). Homooligomeric, wild-type Rdl-receptors expressed in Xenopus oocytes showed the expected electrophysiological properties of GABA-receptors; channels containing the Ala302-Ser mutation, when similarly expressed, were identical with the wild-type ones but had consistently lower sensitivity to dieldrin and PTX, sensitivity to these being reduced about 100-fold. Notably, Lee et al. (1995) reported no detectable specific [3H]-EBOB binding in resistant D. melanogaster strains carrying the Ala302-replacement, despite high specific binding to membranes from susceptible flies, indicating the involvement of Ala302 in EBOB binding to GABA-receptors containing Rdl subunits. Similar results were found for the dieldrin resistance mutation Ala302-glycine found in Drosophila simulans (FfrenchConstant et al., 1993a,b). Furthermore, EBOB blocks chloride ion currents generated by Rdl homomultimers expressed in insect cells and the Ala302-Ser replacement reduces sensitivity to this block 10-fold (Lee et al., 1995). Examination of the Rdl gene from three different insect orders has revealed that in all cases Ala302 is replaced by either a serine or a glycine (less effectively), indicating that this mechanism is universal. The change confers reduced

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sensitivity to PTX, lindane, and TBPS, lower channel conductances, extended channel open times and shorter closed times, and a markedly reduced rate of GABA-induced receptor desensitization. From a simple model to represent binding and allosteric changes, it has been suggested that the preceding mutations are the only ones that can directly weaken cyclodiene binding to the desensitized (antagonistfavored) conformation of the receptor and simultaneously destabilize the antagonist-favored conformation through an allosteric mechanism, resulting in a powerful dualresistance mechanism (Ffrench-Constant et al., 1995; Zhang et al., 1994). In this model, the antagonist associates with the open channel but binds much more tightly when the channel next changes into the desensitized (closed) state, so that this configuration is stabilized. If homomultimers are present in vivo, which may not, however, be the case (Zhang et al., 1995), the preceding mutation could lead to a resistant ion channel with a ring of up to five serines replacing the five alanines in the wild-type Rdl ion channel, which would greatly increase the polarity of this region (5-CH2OH replacing 5-CH3) and considerably alter its affinity for the various toxicants under discussion; even one or two added hydroxyl groups introduced here in a heteromultimer might have a significant effect and also reduce the energy barriers for ion permeation by participating in hydrogen bonding with water (Leonard et al., 1988). Analogies with the nACh receptor are evident and the closed configuration of the mutated chloride channel may remain somewhat ion permeable (Zhang et al., 1994; Revah et al., 1991). Applying sitedirected mutagenesis to explore the human beta 3 homopentamer GABA receptor, which is similar to the insect GABA receptor in insecticide selectivity and specificity, it was demonstrated that the M2-channel lumen facing residues Ala302, Thr306, and Leu309 are significant for NCB binding (Chen et al., 2006).

96.4.3  Molecular Toxicology   of Noncompetitive Chloride   Ionophore Blockers The localization of Ala302 to the PTX/cyclodiene binding site in Rdl prompts some further consideration of the information on the structure–toxicity relationships outlined in earlier sections. The pore cylinder formed by the amino-acid sequences Leu309, Thr306, Leu303, Ala302, Val301 in five adjacent M2 domains (helices) provides a quasicentrosymmetrical lipophilic pocket into which cyclodiene insecticides and other noncompetitive chloride ionophore blockers (NCBs) might fit. Hisano et al. (2007) found that Thr306 has a more significant effect than Ala302 on the binding of EBOB in the pore. The molecular dimensions of cyclodienes, taking into account the known range of allowable molecular substituents in these rather compact

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molecules, are sufficient to block a 5.6-Å pore. If the pore is more or less centrosymmetrical, a cyclodiene molecule could fill the pocket and interact with all of its walls simultaneously because similar binding sites are presented around the lumen even if the arrangement is not a homomultimer. This would explain the observed toxicity of both enantiomeric forms of asymmetric molecules such as heptachlor epoxide; the forms may differ somewhat in toxi­city, however, if the channel is not completely symmetrical, as is found (Miyazaki et al., 1980). Symmetrical molecules should be particularly effective, because they may be able to offer a symmetrically distributed electronegative center (or centers) to similar binding sites on opposite sides of the channel, as in the case of delta-chlordane and isobenzan, each having two symmetrically substituted chlorines on their five-membered rings. These molecules may be viewed as symmetrical composites of the enantiomeric forms of alpha-DH and heptachlor epoxide (HE160), respectively, as suggested in Section 96.3.1 In a hypothetical model (Figure 96.13), for the purpose of structure activity discussion, in which the HCNB moiety of cyclodienes is presumed bound at the synaptic end of the lipophilic pocket so that its gem-dichloro-bridge is presented to the channel wall, then the second electronegative center in, for example, dieldrin, is directed toward the cytoplasmic end of the pocket with its epoxide ring and unchlorinated methano-bridge fixed in an inward direction toward the channel lumen. This “cytoplasmic” end of the molecule is then close to the critical ring of Ala302-methyl groups around the channel lumen, which when replaced by -CH2OH groups inhibits the binding of NCBs. Dieldrin and alpha-endosulfan (extended molecules) appear to provide the limiting acceptable molecular “lengths,” as noted earlier, whereas isobenzan, endrin, and beta-endosulfan are more compact. On this model, it might be argued also that the anti-10-chlorine atom (Figure 96.9 and Table 96.4) of the dichloromethano-bridge of cyclodienes is better accommodated in the lipophilic pocket than the syn-10-chlorine, which may interfere sterically with the large side-chain alkyl groups of the ring of Leu303s that lie toward the synaptic end of the pocket, making this syn-chlorine universally unfavorable for toxicity. The same argument might explain the fourfold increase in dieldrin toxicity effected by removal of the ethylenic chlorines (in DD, Table 96.4), because these chlorines might also interfere sterically in this region. This increase in toxicity is not universal for cyclodienes, however, and the syn-chlorine atom is still present in DD, so that its adverse effect on toxicity is more than offset by removal of the ethylenic chlorine atoms. Further reductive replacement of the syn-10-chlorine atom in DD to give the trichloro-derivative DSD (24, Figure 96.7) reduces toxicity to the level of dieldrin again (Brooks and Mace, 1987). The syn-10-chloro-isomer of DSD is significantly less toxic than DSD or dieldrin, again indicating the greater importance of the 10-anti-chlorine for toxicity. The difference in

S S

O

S

O D

C

S

A H

O B

H

X

P

Y Figure 96.13  Dieldrin (A) is oriented in a hypothetical binding site in or near the chloride ion channel lumen (in the region of Leu303?) with chlorine X (10-anti-chlorine) located in a subsite P that accommodates a bulky substituent. Its epoxide ring then penetrates a three-dimensional region (lower S) near Thr306, which may interact, especially in the closed channel configuration, with the electronegative moieties of various cyclodienes when similarly oriented. If, however, Y (the 10-syn-chlorine) is presented to P, then the epoxide ring cannot so readily interact with zone S (dieldrin orientation B). C indicates the approximate position of the epoxide ring of endrin and also of the sulfur of beta-endosulfan, when either X or Y in these molecules is bound to subsite P; D is the approximate position of the endrin methano-bridge when its 10-anti-chlorine is located in P, corresponding to dieldrin orientation A. Note that a substituted benzene ring and some other extensions are permissible in the upper arrow direction (upper S) toward Leu309 when the bridging system is unchlorinated (see Figure 96.10 and related discussion, Section 96.4.3).

toxicities between syn- and anti-10-monodechloro-isomers is less marked for endrin and beta-endosulfan (the endrinlike isomer) but remains evident for alpha-endosulfan. This observation was discussed (Brooks, 1992) in connection with the insect cross-resistance spectrum for lindane/cyclodienes first noted by Busvine (1964), in which lindane, isobenzan, endrin, and the endosulfan isomers retain measurable toxicity to dieldrin-resistant insects (Brooks and Harrison, 1964a; Busvine, 1964). The first three molecules and beta-endosulfan are rather compact compared with dieldrin and alpha-endosulfan; the latter has been considered to be extended and dieldrin-like (Figure 96.6), but recent structural studies indicate a more complex situation (see below in this section). If the 10-anti-chlorine of dieldrin (X in Figure 96.13) corresponds to the bulky anti-substituent found in PTX and must be presented to an appropriate lipophilic pocket in a binding subsite (P, Figure 96.13) so as to place the epoxide ring in a correct position (in the region of S) with respect to the remainder of the binding site, then the 10-syn-chlorine (Y in Figure 96.13), if similarly presented, cannot place the epoxide ring in the same position. If this latter position is modified in resistance to prevent interaction with the epoxide ring, dieldrin can no longer bind; however, either bridge-chlorine of endrin or beta-endosulfan

Chapter | 96  Interactions with the Gamma-Aminobutyric Acid A-Receptor

can be offered to the bulky substituent binding subsite (P) such that the epoxide ring or sulfite moiety will still be placed in approximately their original positions (near C), still able to interact with the critical site (S), on account of the more compact “cage” shape of these molecules. Consequently, these molecules may still be able to interact to some extent with the binding site that has been modified to exclude binding with dieldrin. On this model, alpha-endosulfan is anomalous in retaining effectiveness, because the same arguments apply as to dieldrin, yet this molecule was actually somewhat more toxic than betaendosulfan to dieldrin-resistant houseflies (Brooks and Harrison, 1964a). Two further observations may be significant, however. First, endosulfan sulfate (4, Figure 96.6) formed in vivo from both endosulfan isomers may be the critical toxicant; it has the same structural (cage) configuration as beta-endosulfan (Forman et al., 1965) and is formed faster from alpha- than from beta-endosulfan in some living organisms. Second, alpha-endosulfan has been reported (Schmidt et al., 1997) to exist in the asymmetrical “twist-chair” conformation, in which the C-O bonds are “trans,” not parallel as usually depicted (Figure 96.6). Molecular models suggest that this twisted configuration may be more flexible, allowing the S5O moiety to occupy several spatial positions between the extremes represented by beta-endosulfan and the extended alpha-structure shown in Figure 96.6. Consequently, the alpha-isomer might be expected to be intrinsically at least as effective as the beta-isomer in terms of their interactions with the resistance-modified binding site, regardless of possible oxidation to the sulfate in vivo. In the case of endrin, the 9-keto-(12-keto-) metabolite is presumed to be the ultimate toxicant in mammals (Hutson et al., 1975) and may contribute to endrin toxicity in insects (Kadous and Matsumura, 1982); notably, this oxidation places a second, additional, electronegative center at D, near the upper subsite S (Figure 96.13), which may improve binding potency toward the resistance-modified binding site. Lindane resembles a very compact cyclodiene and might bind without conflict with a subsite modified for dieldrin resistance or in more than one orientation, and similar arguments apply to isobenzan. Interestingly, isobenzan may be regarded as a “composite” of the HE160 enantiomers, in which an “in plane” oxygen replaces the epoxide rings. Dieldrin resistance normally confers total resistance to HE160 (Brooks and Harrison, 1964a; Busvine, 1964) so the more compact placement of oxygen in isobenzan, combined with a possible increase in binding affinity associated with the symmetrical chlorine substituents, appears to overcome both dieldrin and HE160 resistance to some extent. Notably, diaza-aldrin (3, Table 96.4), in which the unchlorinated double bond is replaced by -N  N- and which is probably converted into its N-oxide (4, Table 96.4) in vivo, is much more toxic than dieldrin to some insects (Busvine, 1964; Soloway, 1965)

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but dieldrin-resistant insects are immune to it. Using the TBPS binding assay in rat brain membranes, it has recently been confirmed that this molecule inhibits the binding competitively and therefore interacts directly with the PTX binding site (Ozoe et al., 1995). Other modifications of the dichloroethylene moiety of dieldrin are acceptable to the binding site, as in photodieldrin (converted in vivo into Klein’s ketone, however; Figure 96.3) and the ketone analogues (14, 15, 16, Figure 96.7) derived from isobenzan. From the analogy with PTX, the dichloroethylene moiety, unchlorinated double bond, and ketone derivatives in these various analogues may correspond to the lactone system of PTX. In lindane, the best superimposition with PTX/cyclodienes is apparently that in which two of the axial chlorines substitute for the lactone ring of PTX and the third axial chlorine provides the equivalent of the bulky trans-isopropenyl group of PTX or anti-10-chlorine of cyclodienes. This last configuration of lindane is favored electrostatically (Calder et al., 1993), although in it the bridge-end chlorines, which complete the pentagonal arrangement of chlorine atoms seen in cyclodienes (Brooks and Mace, 1987), are replaced by hydrogens. Replacement of the bridge-end chlorines by hydrogen reduces the toxicity of some cyclodienes (Soloway, 1965) and, notably, whether bulky alkyl groups in the alpha- and beta-positions of gamma-butyrolactones (cf. PTX) stabilize the open or closed states of the ionophore and hence induce anticonvulsant or convulsant activity depends markedly on the stereochemistry of these substituents relative to the carbonyl group (Holland et al., 1995; Klunk et al., 1983; Peterson et al., 1994). However, the replacement of one bridge-end carbon atom of dieldrin by nitrogen (available for hydrogen bonding) gives azadieldrin (25, Figure 96.11) without great loss in toxicity compared with dieldrin (Gladstone and Wong, 1977), another indication that some modification in this region of cyclodienes is possible. The preceding discussion indicates that numerous modifications of the HCNB moiety retain binding capa­ city to the critical site. In this context, it may be noted that BIDN (34, Figure 96.11), in which the simple norbornene nucleus carries two strongly electron-withdrawing substituents (gem-di-CN and gem-di-CF3), is highly toxic to both mammals and insects (Kölbl et al., 1981). The allowable length compatible with toxicity of the fully chlorinated cyclodienes appears to be restricted, however (see above in this section). TBPS has approximately the same molecular length as cyclodienes such as dieldrin, but there is the question of how the binding of cyclodienes to the critical site relates to that of the extended unchlorinated molecule EBOB or similar molecules in which an aromatic substituent replaces the bulky 4-alkyl group (Palmer et al., 1991a). Superimposition of cyclodienes, aryl-TBOs, aryldithianes, aryl-silatranes (35, Figure 96.11), and PTX by CoMFA (comparative molecular field analysis) (Calder et al., 1993) supports the view that all act at the same or

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overlapping sites. Interesting additional information is available from work on the molecular hybrids of dechlorinated alpha-endosulfan (Figure 96.10) (Ozoe et al., 1993) and insecticidal dioxans (Palmer and Casida, 1995) mentioned previously (Section 96.3.2.2). In this series (Ozoe et al., 1993), the fully chlorinated molecule (6, Figure 96.10) cannot be extended but has the same housefly toxicity as the unchlorinated extended molecule (5, Figure 96.10) related to the compounds reported by Casida. The latter molecule cannot be chlorinated; a single bridge chlorine abolishes its housefly toxicity. Assuming that the hybrid molecule (5) is a rigid form of the corres­ ponding dioxan in which the unchlorinated norbornene moiety serves as the bulky substituent (e.g., t-butyl) (Ozoe et al., 1990) and occupies a spatial region equivalent to that occupied by the HCNB moiety of cyclodienes, it is evident that the combination of substituted phenyl, dioxan, and norbornene moieties can bind to the receptor and afford potentially excellent insect toxicity, as found in 5-t-butyl-2(4-ethynylphenyl)-1,3-dioxane (7, Figure 96.10). However, chlorination may disrupt this binding in the corresponding deschloro-alpha-endosulfan analogue with a 4-cyanophenyl substituent by forcing the “extended” molecule into a position it cannot occupy in the ion channel for steric reasons. The more refined model (Figure 96.14) of the NCB binding site derived from directed mutagenesis of amino-acid residues in the GABA pore channel provides more insights on this problem (Casida and Tomizawa, 2008). Thus, the compact unchlorinated norbornene moiety may not by itself have sufficient binding potency in the lipophilic pore cylinder defined by the Leu309, Thr306, and Ala302 rings, but an added aromatic ring with an appropriate 4-substituent (particularly ethynyl), which, according to the model of Casida and Tomizawa (2008) (Figure 96.14), binds additionally in a region of the channel pore near Leu309, may greatly reinforce the interaction. Conversely, the bulky, fully chlorinated cyclodiene interacts well with the closed configuration of the spherical pocket, but lengthwise extension may be unfavorable in this case because the added benzene ring in the extended HCNB moiety (Figure 96.10) would be forced into steric hindrance with the Leu309 isobutyl groups. Support is given to this hypothesis by the application (Akamatsu et al., 1997) of CoMFA analysis to compare alpha-endosulfan analogues (their series 2) with the hybrid extended norbornene derivatives (series 1) (Ozoe et al., 1993; Figure 96.10) analogous to the 1,3-dioxanes reported by Casida (Figure 96.10). When the simple cyclodienes were closely superimposed on the extended molecules with a 2-(4-cyanophenyl) substituent, the housefly toxi­ city of some of the extended molecules was not well predicted until, in the superposition, the extended molecules were rotated 15° clockwise about their common bond on the norbornene ring junction (C4a–C8a in Table 96.4). This rotation enabled the separate correlation equations for the two series derived by CoMFA on the basis of close

MAMMALIAN BETA-3

A R VA L G I T T V LT

SUBUNIT SEQUENCE RDL WILD TYPE

A R VA L G V T T V LT

RDL RESISTANT MUTANT

A R V S L G V T T V LT

302 306 309 R V L G I T V T

A A

T

L

R V L G I

A

O O O

A A

R V L G I

A N

Q

T

T V

L

T

A

A

N S CF3

T V

T

C1

N

T

L

CF3

NH2 C1

R V L G I

T

T V

L

T

Figure 96.14  Top. Cytoplasmic end of the M2 helix of the human GABAA-receptor 3 -subunit, compared with the corresponding subunit sequences of the Drosophila Wild Type and Drosophila Dieldrin-Resistant mutant (Ala302 Ser-underlined). Bottom. Two of the same five 3-subunit sequences that together define the pore channel of the 3-homopentamer, emphasizing the amino-acid residues exposed to the channel lumen and their proposed binding interactions with EBOB (left) and fipronil (right). (Casida and Tomizawa, 2008).

superimposition to be combined satisfactorily into a single equation representing the housefly toxicities of both series. The two series compared could not, however, be brought together in the same way when the measure of biological activity was the displacement of [35S]-TBPS binding from rat brain membranes. In compound 5 (Figure 96.10), the previous rotation turns the aromatic ring toward the center of the ion channel as modeled in Figure 96.13, away from a region sterically forbidden according to CoMFA, and incidentally may “rock” the unchlorinated norbornene moiety into closer contact with the lipophilic pocket than it can achieve when chlorinated. Conversely, chlorination of the norbornene moiety of these extended molecules would have the reverse effect, forcing the aromatic ring toward the channel wall, into sterically forbidden space. The extending group in TBOs need not necessarily be aromatic, because short alkyl chains with a terminal ethynyl group give insecticidal activity, especially when synergized (Smith et al., 1993) as do 4-ethynylcyclohexyl groups in TBOs and dithianes (Weston et al., 1995). A further interesting feature of the dithianes is their conversion into sulfoxides and sulfones, which doubtless occurs in vivo; the equatorial (trans- with respect to t-butyl; linear) 5-t-butyl-2-(4-Br-phenyl)-1,3-dithiane (8, Figure 96.10; two-substituent is equatorial 4-Br-Ph) is twofold more toxic than the cis- (2-axial; angular) isomer, and the corresponding 4-ethynylphenyl-isomers have equal toxi­city when synergized; for each isomer, conversion into the isomeric monosulfoxides and monosulfone progressively increases toxicity (Palmer and Casida, 1992). The not dissimilar toxicities of the trans- (linear) and cis- (angular) dithianes is intriguing. If, in the pore model (Figure 96.13), the bulky t-butyl substituent is placed in the lipophilic pocket as for the other bulky substituents discussed previously, then the sulfoxides and sulfones from the isomeric dithianes occupy rather different positions in this lipophilic site; only in the linear isomer do these moieties occupy

Chapter | 96  Interactions with the Gamma-Aminobutyric Acid A-Receptor

positions near the two ester oxygens of alpha-endosulfan and the epoxide ring of dieldrin, near Ala302. All of the molecules discussed can be superimposed, and with the cyclodiene molecules oriented in this way, the sulfite moiety of beta-endosulfan (and the SO2 of endosulfan sulfate) lies in the region occupied by the aromatic ring of the angular dithiane isomer. It should be noted, however, that the possibility of cis- to trans-rearrangement exists for the dithianes in vivo (Pulman et al., 1996), in which case, the cis-isomers would merely be precursors of the linear (trans-) molecules, which are more readily accommodated in the model. Among the heterocyclic compounds of recent interest that are believed to act by blocking the GABA-gated chlor­ ide ionophore, the experimental spirosultam LY 219048 (33, Figure 96.11) contains an obvious lipophilic bulky substituent in the form of the cyclohexane ring, in analogy with the compounds discussed previously. Probably because this ring may be susceptible to metabolic attack, the compound is not very toxic to insects but its toxicity to mice is similar to that of endrin (Bloomquist et al., 1993). Other insecticidal compounds such as the phenylpyrazoles (e.g., fipronil; 27, Figure 96.11) have little obvious structural similarity to the compounds discussed in previous sections yet inhibit [3H]-EBOB binding to housefly head membranes and are believed to act at the PCCA/PTX binding site (Cole et al., 1993). Some other compounds containing the common 2,6dichloro-4-trifluoromethylphenyl moiety combined with a pyrimidinone (e.g., 32, Figure 96.11) and other small heterocyclic rings were described by Whittle et al. (1995), but these lacked the broad-spectrum insecticidal activity shown by fipronil, which was introduced in 1993 (Colliot et al., 1992) and now has a wide range of applications for crop protection by foliar, soil, or seed treatment. Its mammalian toxicity is generally moderate (rat acute oral LD50, 97 mg/ kg; mouse acute oral LD50, 95 mg/kg) and it is readily converted into degradation products, including the corresponding sulfone, in the environment (Tomlin, 1997). The substitution pattern in the phenyl and hetero­cyclic rings of these compounds places these ring planes at right angles (Whittle et al., 1995) so that the molecules are “space filling” in these two planes and are relatively rigid, with the pyrazole ring skewed with respect to the benzene ring because of the pyramidal linking nitrogen. The substituents (-NH2, -CN, -SOCF3) are attached to double bonds and are fixed in the plane of the pyrazole ring, although they can rotate about their attaching bonds, which may be particularly important for the critical -SOCF3 group. This group confers insect (housefly LD50, 0.3 g/g) and vertebrate (mouse ip LD50, 30 mg/kg) toxicity, even in the analogous molecule lacking both the -NH2 and -CN substituents, which, however, clearly optimize the binding properties of fipronil. In these molecules, the sulfur atom requires bioactivation through sulfoxide/sulfone formation to confer toxicity. In the related insecticide ethiprole, used in rice paddy fields, -SOCH2CH3 replaces -SOCF3.

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How do the phenylpyrazoles bind to the chloride iono­ phore in relation to the other molecules discussed previously? They inhibit EBOB binding in housefly head membranes noncompetitively, which may involve irreversible or slowly reversible inhibition or action at an allosteric but coupled site (Cole et al., 1993). Whittle et al. (1995) explored a series of compounds, including pyrimidinone 32 (Figure 96.11) on the basis of their commonality of dipole direction with phenylpyrazoles; the positive end of the dipole lies toward the benzene ring in active compounds. If this ring, with its strongly electron-withdrawing -CF3 moiety, is placed in the channel pore binding site so that the -CF3 is in a similar position to that occupied by the 4-ethynyl group of EBOB relative to Leu309 as in Figure 96.14 (Casida and Tomizawa, 2008), then the benzene ring with its chlorine atoms and the pyrazole ring with its substituents are well placed to bind with channel components in this quasi-centrosymmetrical ionophore and, notably, the -SOCF3 moiety is then placed near Ala302 in a region occupied by the endosulfan ester oxygens, the oxygens of TBOs, the oxygens or sulfurs of dioxans and oxidized trans- (linear) dithianes, when these molecules are located in the lipophilic region near Leu303. Other studies on the interactions between fipronil analogues and a docking model of the beta 3-homopentamer were reported by Alam et al. (2007). These molecules might penetrate the chloride ionophore from either the extracellular (synaptic) mouth or the cytoplasmic mouth or by first penetrating the lipid bilayer and then entering the channel laterally. Current understanding of the nACh-receptor (nAChR) ionophore, which has analogies with the GABAA ionophore, may be relevant. According to Unwin (1995), the M2 cylinders (alpha-helices) (Figure 96.12) are kinked inward in the region of a ring of leucines when the ionophore is closed; in the open configuration, the cylinders are twisted laterally, moving the leucine side chains away from the pore. A refined model of the membrane-associated Torpedo acetylcholine receptor at 4 Angstrom resolution has been obtained by electron microscopy (Unwin, 2005). By analogy, a cyclodiene (for example) may enter a similar open conformation in the chloride ionophore and then become tightly bound when the pore reverts to the closed conformation, as suggested by Ffrench-Constant et al. (1995). Thus, in the case of the GABAA-receptor, multiple hydrogen bonding plus hydrophobic interactions in the Ala302 to Leu309 region must have a significant role in the action of the cage convulsants in blocking the chloride ionophore (Figure 96.14). Here we recall that the hybrid (unchlorinated) molecule 4 (Figure 96.10) combines the bulky norbornene (hydrophobic binding) and dioxan moieties with the additional binding capacity evidently conferred by the benzene ring with its electronegative 4-ethynyl substituent; it is more toxic (when oxidative metabolism is inhibited) than many fully chlorinated cyclodienes. Nakanishi et al. (1997) provide an interesting discussion on the question of mode of channel entry, based

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on research on philanthotoxins (PhTXs) binding to the nAChR. In particular, they noted that an n-butyl side chain introduced into the hydrophilic polyamine chain of PhTX433 increases potency eightfold. The n-butyl moiety is so placed as to reach the area of Leu251 when the molecule is inserted linearly into the AChR ionophore from the cytoplasmic end, and the authors speculate that the increased potency might result from additional hydrophobic binding between the n-butyl side chain and the alkyl groups of Leu251; potency decreases below that of the parent PhTX433 when this side chain is placed in other positions along the polyamine chain.

Conclusion Some 45 years after serious toxicological research on the PCCA insecticides began, there is now a broad understanding of their mode of action and, due to rapid progress in the application of molecular biology, the tantalizing mechanism of insect resistance to them has been illuminated. Furthermore, new structural classes of chemicals have now emerged that appear to act in the same way. Although many chemicals found to interact with the picrotoxinin binding site in the GABAA-receptor chloride ionophore are highly toxic to both mammals and insects, binding studies with radioligands have indicated differences between their GABA-receptors that offer the prospect of selective insect toxicity involving this target. The commercially successful insecticide fipronil appears to fulfill these expectations and other new chemicals are likely to follow. There is also the possibility of “building in” selectivity by using the “propesticide” approach, which exploits differences between insects and nontarget organisms in their biotransformation routes for chemically derivatized toxicants and has been applied successfully to alleviate the mammalian toxicity of other classes of insecticides. Meanwhile, many questions remain to be answered that are of fundamental importance in understanding the molecular action of neurotransmitters and insecticides on ion channels. The intense interest in this subject, stimulated for many years by the resistance problem and more recently by rapid advances in molecular biology (Gammon and Brooks, 2007), will ensure a prominent use for the PCCA insecticides and the newer chemicals with related actions, as tools in these explorations.

References Abalis, I. M., Eldefrawi, M. E., and Eldefrawi, A. T. (1985). High affinity stereospecific binding of cyclodiene insecticides and gamma-hexachlorocyclohexane to gamma-aminobutyric acid receptors. Pestic. Biochem. Physiol. 24, 95–102. Akamatsu, M., Ozoe, Y., Fujita, T., Mochida, K., Nakamura, T., and Matsumura, F. (1997). Sites of action of noncompetitive GABA

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antagonists in houseflies and rats: Three dimensional QSAR analysis. Pestic. Sci. 49, 319–332. Akkermans, L. M. A., van den Bercken, J., van der Zalm, J. M., and van Stranten, H. W. M. (1974). Effects of dieldril and some of its metab­ olites on synaptic transmission in the frog motor end plate. Pestic. Biochem. Physiol. 4, 313–324. Akkermans, L. M. A., van den Bercken, J., and van der Zalm, J. M. (1975a). Effects of aldrin-transdiol on neuromuscular facilitation and depression. Eur. J. Pharmacol. 31, 166–175. Akkermans, L. M. A., van den Bercken, J., and Versluijs-Helder, M. (1975b). Excitatory and depressant effects of dieldrin and aldrintransdiol on the spinal cord of the toad (Xenopus laevis). Eur. J. Pharmacol. 34, 133–142. Alam, M. S., Huang, J., Ozoe, F., Matsumura, F., and Ozoe, Y. (2007). Synthesis, 3D-QSAR, and docking studies of 1-phenyl-1,2,3 triazoles as selective antagonists for beta3 over alpha1beta2gamma2 GABA receptors. Bioorg. Med. Chem. 15, 5090–5104. Balba, H. M., and Saha, J. G. (1978). Studies on the distribution, excretion and metabolism of alpha- and gamma-isomers of [14C] chlordane in rabbits. J. Environ. Sci. Health, B 13, 211–233. Baldwin, M. K. (1971). “The Metabolism of the Chlorinated Insecticides Aldrin, Dieldrin, Endrin and Isodrin.” Ph.D. Thesis, University of Surrey, Guildford, UK. Baldwin, M. K., and Robinson, J. (1969). Metabolism in the rat of the photoisomerisation product of dieldrin. Nat. (Lond.) 224, 283–284. Baldwin, M. K., Robinson, J., and Parke, D. V. (1972). A comparison of the metabolism of HEOD (Dieldrin) in the CF1 mouse with that in the CFE rat. Food Cosmet. Toxicol. 10, 333–351. Barnes, W., and Ware, G. W. (1965). The absorption and metabolism of C14-endosulfan in the housefly. J. Econ. Entomol. 58, 286–291. Bedford, C. T. (1974). Von Baeyer/IUPAC names and abbreviated chemical names of metabolites and artifacts of aldrin (HHDN), dieldrin (HEOD) and endrin. Pestic. Sci. 5, 473–489. Bedford, C. T., Harrod, R. K., Hoadley, E. C., and Hutson, D. H. (1975a). The metabolic fate of endrin in the rabbit. Xenobiotica 5, 485–500. Bedford, C. T., Hutson, D. A., and Natoff, I. L. (1975b). The acute toxi­ city of endrin and its metabolites to rats. Toxicol. Appl. Pharmacol. 33, 115–121. Bender, H. (1935). Benzene hexachloride and other chlorinated carbocyclics. U.S. Patent 2010841. Bloomquist, J. R., and Shankland, D. L. (1983). The mode of action and neurotoxicity of mirex, chlordecone and four hydrogenated mirex analogues. Pestic. Biochem. Physiol. 19, 235–242. Bloomquist, J. R., Jackson, J. L., Karr, L. L., Ferguson, H. J., and Gajewski, R. P. (1993). Spirosultam LY 219048: A new chemical class of insecticide acting upon the GABA receptor/chloride ionophore complex. Pestic. Sci. 39, 195–202. Boddy, K. L., Briggs, G. G., Harrison, R. P., Jones, T. H., O’Mahony, M. J., Marlow, I. D., Roberts, B. G., Bardsley, R., and Reid, J. (1996). The synthesis and insecticidal activity of 2-aryl-1,2,3-triazoles. Pestic. Sci. 48, 189–196. Brimfield, A. A., and Street, J. C. (1979). Mammalian biotransformation of chlordane: in vivo and primary hepatic comparisons. Ann. N.Y. Acad. Sci. 320, 247–256. Brimfield, A. A., Street, J. C., Futsell, J., and Chatfield, D. A. (1978). Identification of products arising from the metabolism of cis- and trans-chloruane in rat liver microsomcs in vitro: Outline of possible metabolic pathway. Pestic. Biochem. Physiol. 9, 84–95. Brooks, G. T. (1960). Mechanisms of resistance of the adult housefly (M. domestica) to cyclodiene insecticides. Nat. (Lond.) 186, 96–98.

Chapter | 96  Interactions with the Gamma-Aminobutyric Acid A-Receptor

Brooks, G. T. (1966). Progress in metabolic studies of the cyclodiene insecticides and its relevance to structure–activity correlations. World Rev. Pest Control 5, 62–84. Brooks, G. T. (1969). The metabolism of diene-organochlorine (cyclodiene) insecticides. Residue Rev. 27, 81–138. Brooks, G. T. (1973). The design of insecticidal chlorohydrocarbon derivatives. In “Drug Design” (E. J. Ariens, ed.), Vol. 4, pp. 379–444. Academic Press, New York. Brooks, G. T. (1974). “Chlorinated Insecticides,” pp. 87–98. CRC Press, Cleveland. Brooks, G. T. (1975). The insect toxicities of biodegradable derivatives of chlorinated norbornenes. In “Proceedings of the Eighth British Insecticide and Fungicide Conference,” Vol. 2, pp. 381–387. British Crop Protection Council, Farnham, Surrey, UK. Brooks, G. T. (1977). Action and inaction of certain non-anticholinesterase insecticides. In “Proceedings of the 1977 British Crop Protection Conference – Pests and Diseases,” Vol. 3, pp. 731–740. British Crop Protection Council, Farnham, Surrey, UK. Brooks, G. T. (1980). The preparation of some reductively dechlorinated analogues of dieldrin, endosulfan and isobenzan. J. Pestic. Sci. (Tokyo) 5, 565–574. Brooks, G. T. (1985). The preparation of reductively dechlorinated analogues of endrin and some other cyclodiene insecticides. J. Pestic. Sci. (Tokyo) 10, 241–245. Brooks, G. T. (1992). Progress in structure–activity studies on cage convulsants and related GABA receptor chloride ionophore antagonists. In “Insecticides: Mechanism of Action and Resistance” (D. Otto and B. Weber, eds.), pp. 237–242. Intercept, Andover, UK. Brooks, G. T., and Harrison, A. (1963). Relations between structure, metabolism and toxicity of the “cyclodiene” insecticides. Nat. (Lond.) 198, 1169–1171. Brooks, G. T., and Harrison, A. (1964a). The effect of pyrethrin synergists, especially sesamex, on the insecticidal potency of hexachlorocyclopentadiene derivatives (“cyclodiene” insecticides) in the adult housefly, M. domestica L. Biochem. Pharmacol. 13, 827–840. Brooks, G. T., and Harrison, A. (1964b). The metabolism of some cyclodiene insecticides in relation to dieldrin resistance in the adult housefly, M. domestica L. J. Insect Physiol. 10, 633–641. Brooks, G. T., and Harrison, A. (1965). Structure–activity relationships among insecticidal compounds derived from chlordene. Nat. (Lond.) 205, 1031–1032. Brooks, G. T., and Harrison, A. (1967a). The metabolism of dihydrochlordene and related compounds by housefly (M. domestica L.) and pig liver microsomes. Life Sci. 6, 681–689. Brooks, G. T., and Harrison, A. (1967b). The toxicity of alpha-dihydroheptachlor and related compounds to the housefly (M. domestica L.) and their metabolism by housefly and pig liver microsomes. Life Sci. 6, 1439–1448. Brooks, G. T., and Harrison, A. (1969a). The oxidative metabolism of aldrin and dihydroaldrin by houseflies, housefly microsomes and pig liver microsomes and the effect of inhibitors. Biochem. Pharmacol. 18, 557–568. Brooks, G. T., and Harrison, A. (1969b). Hydration of HEOD (dieldrin) and the heptachlor epoxides by microsomes from the livers of pigs and rabbits. Bull. Environ. Contam. Toxicol. 4, 352–361. Brooks, G. T., and Mace, D. W. (1987). Toxicity and mode of action of reductively dechlorinated cyclodiene insecticide analogues on houseflies (M. domestica L.) and other diptera. Pestic. Sci. 21, 129–142. Brooks, G. T., Lewis, S. E., and Harrison, A. (1968). Selective metabol­ ism of cyclodiene enantiomers by pig liver microsomal enzymes. Nat. (Lond.) 220, 1034–1035.

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Reddy, G., and Khan, M. A. Q. (1977). Metabolism of [14C]-photodieldrin in house flies. J. Agric. Food Chem. 25, 25–28. Revah, F., Bertrand, D., Galzi, J. L., Devillers-Theiry, A., Mulle, C., Hussy, N., Bertrand, S., Ballivet, M., and Changeux, J. P. (1991). Mutations in the channel domain alter desensitisation of a neuronal nicotine receptor. Nat. (Lond.) 353, 846–849. Richardson, A., Baldwin, M. K., and Robinson, J. (1968). Metabolites of dieldrin in the urine and faeces of rats. Chem. Ind. (Lond.), 588–590. Ryan, K. J., and Engel, L. L. (1957). Hydroxylation of steroids at carbon21. J. Biol. Chem. 225, 103–114. Saleh, M. A., Skinner, R. F., and Casida, J. E. (1979). Comparative metabol­ ism of 2,2,5-endo,6-exo,8,9,10-heptachloronorbornene and toxaphene in six mammalian species and chickens. J. Agric. Food Chem. 27, 731–737. Schmidt, W. F., Hapeman, C. J., Fettinger, J. C., Rice, C. P., and Bilboulian, S. (1997). Structure and asymmetry in the isomeric conversion of betato alpha-endosulfan. J. Agric. Food Chem. 45, 1023–1026. Schofield, P. R., Darlison, M. G., Fujita, N., Burl, D. R., Stephenson, F. A., Rodriguez, H., Rhee, L. M., Ramachandran, J., Reale, V., Glencorse, T. A., Seeburg, P. H., and Barnard, E. A. (1987). Sequence and functional expression of the GABAA-receptor shows a ligandgated receptor super-family. Nat. (Lond.) 328, 221–227. Schroeder, M. E., Shankland, D. E., and Hollingworth, R. M. (1977). The effects of dieldrin and isomeric diols on synaptic transmission in the American cockroach and their relevance to the dieldrin poisoning syndrome. Pestic. Biochem. Physiol. 7, 403–415. Shankland, D. L. (1982). Neurotoxic action of chlorinated hydrocarbon insecticides. Neurobehav. Toxicol. Teratol. 4, 805–811. Shankland, D. L., and Schroeder, M. E. (1973). Pharmacological evidence for a discrete neurotoxic action of dieldrin (HEOD) in the American cockroach, Perplaneta americana (L.). Pestic. Biochem. Physiol. 3, 77–86. Slade, M., Brooks, G. T., Hetnarski, H., and Wilkinson, C. F. (1975). Inhibition of the enzymatic hydration of the epoxide HEOM in insects. Pestic. Biochem. Physiol. 5, 35–46. Slade, R. E. (1945). The gamma-isomer of hexachlorocyclohexane (gammexane). Chem. Ind. (Lond.) 40, 314–319. Smith, A. G. (1991). Chlorinated hydrocarbon insecticides. In “Handbook of Pesticide Toxicology” (W. J. Hayes Jr. and E. R. Laws Jr., eds.), Vol. 2, pp. 731–915. Academic Press, New York. Smith, I. H., Budd, T. C., Sills, J. H., and Casida, J. E. (1993). Insecticidal 1-(alkynyl alkyl)-3-cyano-2,6,7-trioxabicyclo[2.2.1] octanes. J. Agric. Food Chem. 41, 1114–1117. Soloway, S. B. (1965). Correlation between biological activity and molecular structure of the cyclodiene insecticides. Adv. Pest Control Res. 6, 85–126. Sternburg, J., Kearns, C. W., and Moorefield, H. (1954). DDT-dehydrochlorinase, an enzyme found in DDT-resistant flies. J. Agric. Food Chem. 2, 1125. Street, J. C., and Blau, S. E. (1972). Oxychlordane. Accumulation in rat adipose tissue on feeding chlordane isomers or technical chlordane. J. Agric. Food Chem. 20, 395–397. Sun, Y. P., and Johnson, E. R. (1960). Synergistic and antagonistic actions of insecticide – synergist combinations and their mode of action. J. Agric. Food Chem. 8, 261–266. Takeuchi, A., and Takeuchi, N. (1966). On the permeability of the presynaptic terminal of the crayfish neuromuscular junction during synaptic inhibition and the action of GABA. J. Physiol. 183, 433–449. Takeuchi, A., and Takeuchi, N. (1972). Actions of transmitter substances on the neuromuscular functions of vertebrates and invertebrates. Adv. Biophys. 3, 45–95.

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Tanaka, K., Scott, J. G., and Matsumura, F. (1984). Picrotoxinin receptor in the central nervous system of the American cockroach: Its role in the action of cyclodiene-type molecules. Pestic. Biochem. Physiol. 22, 117–127. Tashiro, S., and Matsumura, F. (1977). Metabolic routes of cis- and transchlordane in rats. J. Agric. Food Chem. 25, 872–880. Tomlin, C. D. S., (ed.) (1997). “Pesticidehonval,” 4th ed., pp. 545–547. British Crop Protection Council, Farnham, Surrey, UK. Turner, W. V., Engel, J. L., and Casida, J. E. (1977). Toxaphene components and related compounds: preparation and toxicity of some hepta-, octa-and nonachloronorbornanes, hexa- and heptachlorobornenes, and a hexachlorobornadiene. J. Agric. Food Chem. 25, 1394–1401. Uchida, M., Fujita, T., Kurihara, N., and Nakajima, M. (1978). Toxicities of gamma-BHC and related compounds. In “Pesticide and Venom Neurotoxicity” (D. L. Shankland, R. M. Hollingworth, and T. Smyth Jr., eds.), pp. 133–151. Plenum, New York. Unwin, N. (1995). Acetylcholine receptor channel imaged in the open state. Nat. (Lond.) 373, 37–43. Unwin, N. (2005). Refined structure of the nicotinic acetylcholine receptor at 4 A resolution. J. Mol. Biol. 346, 967–989. Von Keyserlingk, H. C., and Willis, R. J. (1992). The Gaba-activated chloride channel in insects as target for insecticide action – A-physiological study. In “Insecticides: Mechansim of Action and Resistance” (D. Otto and B. Weber, eds.), pp. 205–236. Intercept, Andover, UK. Wacher, V. J., Toia, R. F., and Casida, J. E. (1992). 2-Aryl-5-tert-butyl1,3-dithianes and their S-oxidation products: Structure–activity relationships of potent insecticides acting at the GABA-gated chloride channel. J. Agric. Food Chem. 40, 497–505. Wang, C. M., Narahashi, T., and Yamada, M. (1971). The neurotoxic action of dieldrin and its derivatives in the cockroach. Pestic. Biochem. Physiol. 1, 84–91. Weil, E. D., Colson, J. G., Hoch, P. E., and Gruber, R. H. (1969). Toxic chlorinated methanoisobenzofuran derivatives. J. Heterocyc. Chem. 6, 643–649. Weston, J. B., Larkin, J. P., Pulman, D. A., Holden, I., and Casida, J. E. (1995). Insecticidal isomers of 4-tert-butyl-1-(4-ethynylcyclohexyl)2,6,7-trioxabicyclo[2.2.1] octane and 5-tert-butyl-2-(4-ethynylcyclohexyl)-1,3-dithiane. Pestic. Sci. 44, 69–74. Whittle, A. J., Fitzjohn, S., Mullier, G., Pearson, D. P. J., Perrior, T. R., Taylor, R., and Salmon, R. (1995). The use of computer-generated electrostatic surface maps for the design of new GABA-ergic insecticides. Pestic. Sci. 44, 29–31. Wong, D. T., and Terriere, L. C. (1965). Epoxidation of aldrin, isodrin and heptachlor by rat liver microsomes. Biochem. Pharmacol. 14, 375–377. Yarbrough, J. D., Grimley, J. M., Karl, P. I., Chambers, J. E., Case, R. S., and Alley, E. G. (1983). Tissue disposition, metabolism and excretion of cis-and trans-5,10-dihydrogen mirex. Drug Metab. Dispos. 11, 611–614. Zhang, H.-G., Ffrench-Constant, R. H., and Jackson, M. B. (1994). A unique amino acid of the Drosophila GABA-receptor with influence on drug sensitivity by two mechanisms. J. Physiol. 479, 65–75. Zhang, H.-G., Lee, H.-J., Rocheleau, T., Ffrench-Constant, R. H., and Jackson, M. B. (1995). Subunit composition determines picrotoxinin and bicuculline sensitivity of Drosophila GABA-receptors. Mol. Pharmacol. 48, 835–840.

Chapter 97

The Role of P-glycoprotein in Preventing Developmental and Neurotoxicity: Avermectins – A Case Study Jim Stevens1, Charles B. Breckenridge2 and Jayne Wright3 1

Wake Forest University School of Medicine, Winston-Salem, North Carolina Syngenta Crop Protection, Inc., Greensboro, North Carolina 3 Syngenta Crop Protection, Inc., Bracknell, United Kingdom 2

97.1  Introduction The avermectins are macrocyclic lactones isolated from the fermentation broth of the soil actinomycete, Streptomyces avermitilis. Included in this avermectin group are abam­ ectin and emamectin benzoate, which are used as insecti­ cides, and ivermectin, which is used for parasite control in human and veterinary medicine. Because the avermec­ tins act as GABAA receptor agonists in vertebrates, their general safety in mammals depends, in part, on an intact blood–brain barrier in juvenile and adult animals and an intact blood–placental barrier in utero. Inherent in the integrity of these barriers is P-glycoprotein, which is an efflux transporter. Intact P-glycoprotein barriers are present in human adults (male and female, pregnant and nonpreg­ nant), newborns, and children. This fact is supported by the safe use of 0.15–0.20 mg/kg of ivermectin in a genetically varied population throughout the world, including thou­ sands of pregnant women, of over 50 million people (FAO/ WHO, 1997). However, there is also experimental evidence that certain laboratory animals, such as genetically poly­ morphic CF-1 mice and the collie dog, do not possess intact P-glycoprotein blood–brain barrier (Lankas et al., 1997; Schinkel et al., 1994). Further, it has been demon­ strated that the P-glycoprotein blood–brain barrier is not fully developed in the neonatal CD-1 mouse (Yavanhxay et al., 2005), the neonatal FVB wild type mouse (Tsai et al., 2002), the neonatal Sprague-Dawley rat (Yavanhxay et al., 2004), or the Wistar rat (Matsuoka et al., 1999). Further, the timetable for the development of the rodent ­placental– P-glycoprotein barrier differs from that of humans, ­placing the developing fetus in a vulnerable position during Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

­organogenesis for developmental effects (Novotna et al., 2004; Aleksunes et al., 2008). Unfortunately, the rodent models have used and will continue to be used as a sur­ rogate to the human to establish a hazard profile for the mectins. This practice has resulted in very low NOELs (no-observed-effect levels) for neurotoxicity and devel­ opmental endpoint as the tests were consulting using the CF-1 polymorphic mouse as the animal model. The World Health Organization’s Joint Meeting on Pesticide Residues (JMPR) and the U.S. Environmental Protection Agency have concluded that the CF-1 polymorphic mouse is not an appropriate model for human risk assessment for the aver­ mectins; the JMPR has recognized the flawed nature of the rat pup model for testing these avermectins. However, a clear appreciation of the inappropriateness of using haz­ ard data for the mectins derived from these standard ani­ mal models during stages developmental of P-glycoprotein ontogeny for human risk assessment is yet to be fully appreciated.

97.2  Chemistry and formulations Abamectin belongs to a general class of closely related macrocyclic lactones either produced directly by the acti­ nomycete Streptomyces avermitilis or generated through semisynthetic modifications (Fisher and Mrozik, 1989). The structure for the natural avermectins is given in Figure 97.1. The basic structural motif of the avermectins is evident in the natural product avermectin B1a, which is the principal constituent of the insecticide abamectin. As used in pesti­ cides, abamectin consists of 80% or more of avermec­ tin B1a and 20% or less of avermectin B1b and is called 2093

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2094

H3C

O

HO H3C

4’’ H3C

O

O

O 4’ H3C

CH3 O

22

H

O

23

O

CH3 CH3

25 O

H3C

R 26

H

O

H

O

OH H 8

2

O

5 H

CH3

OR

Figure 97.1  Structure of the natural avermectins.

OCH3 HO OCH3 CH3

O

O CH3 CH3

O

CH3

H

O

O

CH3 O R

CH3 (i) R = -CH2CH3 (avermectin B 1a)

H

H

O

O

OH

(ii) R = -CH3 (avermectin B 1b)

H

O H

CH3 OH

Figure 97.2  Structures of ivermectin and abamectin.

a­ vermectin B1 (Fisher and Mrozik, 1989). Their structures are shown in Figure 97.2. Chemical modification of aver­ mectin B1a has yielded a number of semisynthetic mater­ ials. Emamectin (4-epimethylamino-4-deoxyavermectin B1a) benzoate is shown in Figure 97.3. Emamectin and

ivermectin differ from avermectin B1 by having only a sin­ gle bond at the C22C23 position (instead of a double bond). The major manufacturers, trade names, and formulations for abamectin, emamectin benzoate, and ivermectin are given in Table 97.1.

Chapter | 97  The Role of P-glycoprotein in Preventing Developmental and Neurotoxicity

2095

O

CH3 H

H N

O

O

CH3

O

H3C H3C

H

CH3

H3C

H O

O

O O

C H

C

H3C

H

CH3 H

CH3

CH2

H

O

CH3

H

H

OH

O

O

OH

H

H HO

H

emamectin B1a benzoate (major component) O CH3

O

CH3 H

H N

O

O

CH3

O

H3C H3C

H O

O C

H

CH3

H3C

CH3

H

O

O

CH3

H

O

C H

H3C

CH3 H

H

H

OH emamectin B1b benzoate (minor component)

O

O

OH

H

O H HO

CH3 H

Figure 97.3  Structure of emamectin benzoate.

97.3  Uses Abamectin and emamectin benzoate (Syngenta) are applied foliarly. They penetrate the leaves, and breakdown rapidly in sunlight; all of these features favor their use in inte­ grated pest management (Bloomquist, 1999). Abamectin is used primarily to control mites, and emamectin benzoate is

used to control lepidopterian species in vegetable, cotton, and tobacco. Ivermectin is used as an anthelmintic in the treatment of infection of intestinal threadworm, river blindness (onchocerciasis), and lymphatic filariasis. Its uses in vet­ erinary medicine have been as anthelmintic and antipara­ sitic agents, including treatment of heartworm, hookworm,

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Table 97.1  The Major Manufacturers, Trade Names, and Formulations for Abamectin, Emamectin Benzoate, and Ivermectin Chemical Abamectin

Emamectin benzoate

Ivermectin

Trade names

Major manufacturers

Formulation 0.15 lb/gal

Agri-mek

Syngenta

Zephyr

Syngenta

0.15 lb/gal

Avid

Syngenta

0.15 Emulsifiable concentrate

Proclaim

Syngenta

5% granule

Denim

Syngenta

0.16 lb/gal

Ivomec

Merck & Co.

Tablets

Mectizan

Merck & Co.

Injectables

Stromectal

Merck & Co.

3-mg Tablets

threadworm, and whipworm (FAO/WHO, 1992, 1993; Greene et al., 1989).

97.4  Mode of action of the avermectins The avermectins were first demonstrated to possess anthel­ mintic activity in 1979 (Burg et al., 1979). The avermectins act as chloride channel blockers, causing hyperexcitability and convulsions. Arena et al. (1995) demonstrated in insects that stimulation of glutamate (inhibitory) chloride channels is the most sensitive target site for the avermectins. The glutamate-gated chloride channels of insect and nematode skeletal muscle are especially important as they mediate avermectin-induced muscle paralysis in these organisms. These effects are mediated via a specific, high-affinity (1010 M) binding site (Turner and Schaeffer, 1989). In vertebrates, the mectins bind to the receptor for the inhibitory neurotransmitter -aminobutyric acid (GABA); see Figure 97.4. The avermectins open the GABAA recep­ tor chloride channel by binding to the GABA recognition site (receptor protein) and act as partial agonists (Abalis et al., 1986). Chloride ions then flow into the postsynap­ tic neuron. This chloride permeability increase can signifi­ cantly hyperpolarize (make more negative) the membrane potential, which has a dampening effect on nerve impulse firing. There is also a reversible dose-dependent increase in chloride ion permeability in the presence of very low doses of avermectins. In GABA-insensitive neurons with no inhibitory inner­ vation, the avermectins induce an irreversible increase in chloride ion conductance by interacting with voltagedependent chloride channels. Avermectin intoxication in mammals begins with hyperexcitability, tremors, and

i­ncoordination and later develops into ataxia and coma-like sedation. This is similar to the mode of action of ethanol and barbiturates (Eldefawi and Eldefawi, 1987) and ben­ zodiazepine sedatives (Williams and Yarbrough, 1979). However, the avermectins are less specific in their action and can affect a variety of other ligand- and voltage-gated chloride channels.

97.5  Hazard identification and dose response As previously indicated, in vertebrates, the avermectins increase membrane permeability to chloride ions and act as GABAA agonists; this is similar to the mode of action of the benzodiazepine sedatives (Turner and Schaeffer, 1989). Their toxicity reflects this mode of action in overdose scen­arios and in animal models with compromised mdr1a P-glycoprotein barriers. The acute toxicology profiles for ivermectin and abamectin (U.S. EPA, 1999a; Lankas and Gordon, 1989) and emamectin benzoate (U.S. EPA, 1999b) are shown in Table 97.2. These three avermectin-derived materials responded quite similarly in the different laboratory models. A com­ parison of the response of ivermectin and abamectin in the monkey as well as the response noted in the human with ivermectin is presented in Table 97.3. Signs of overdosing noted at 24 mg/kg of ivermectin or abamectin in the monkey were the same as observed in the human overdose at approximately 9 mg/kg. These signs were essentially identical to those observed in 10 adults who accidentally ingested tablets or solutions intended for veterinarian use (Greene, 1991). Subchronic dietary exposure to the three avermec­ tins in the rat, mouse, and dog yielded similar results,

Chapter | 97  The Role of P-glycoprotein in Preventing Developmental and Neurotoxicity

2097

OCH3 HO

OCH3 H3C

O H

O H

H3C

CH3 O

O

O

CH3

H

CH3

H

O H

H3C

R

H

H

H

O

O

OH

H

O CH3

H OH

Figure 97.4  a) Insect NERVE Junction (or Synapse)-inhibitory nerve transmission. b) Mode of action of the mectins.

Table 97.2  Acute Oral Toxicity Studies on Ivermectin, Abamectina and Emamectin Benzoateb LD50 (mg/kg)

Species Ivermectin

Abamectin

Emamectin

SD rat

50

11

76–88

SD rat, neonatalc

2

1.5

—

CD-1 mouse

—

220

107–120

d

CF-1 mouse

25

14–24

22–31

Beagle dogs

80

—

—

Rhesus monkeys

24

24

—

a

Lankas and Gordon, 1989; EPA, 1999a. US. PA, 1999b. c P-glycoprotein-deficient blood–brain barrier is seen in neonatal rats (Lankas et al., 1989). d CF-1 mice tested were polymorphic for P-glycoprotein (Umbenhauer et al., 1997). b

as shown in Table 97.4. Although there are slight differ­ ences between the NOELs for abamectin and emamectin in the CD-1 mouse (probably the result of dose selection), the responses noted in the rat and dog were similar for

the three ­ avermectins. The avermectins are equally well ­tolerated following chronic dietary administration, as shown in Table 97.5. The NOELs found in the chronic dog and mouse onco­ genicity studies were comparable for ivermectin, abam­ ectin, and emamectin. The NOEL in the chronic rat study was higher for abamectin than for emamectin or ivermec­ tin. This apparent difference between abamectin and ema­ mectin was most likely due to differences in dose selection between the studies. The avermectins are not genotoxic, as has been demon­ strated in a variety of standard tests for mutagenicity, clas­ togenicity, and unscheduled deoxyribonucleic acid (DNA) synthesis, as presented in Table 97.6. The maternotox­ icity and developmental/fetotoxicity NOELs and lowestobserved-effect levels (LOELs) for these three avermectins are shown in Table 97.7. In the CF-1 mouse, SD rat, and rabbit, developmental toxicity studies with ivermectin, cleft palate and clubbed feet (rabbit only) were observed at maternally toxic doses (Lankas and Gordon, 1989). Similar findings were noted in the CF-1 mouse and rabbit studies with abamectin. Neither of these effects was noted with emamectin (U.S. EPA, 1999b). Sedation was observed in overdosed rabbit dams. Severe neurotoxicity (tremors, convulsion, and coma) was observed in some of the polymorphic CF-1 mice with a compromised blood–brain barrier and blood–placental bar­ rier (Umbenhauer et al., 1997). These effects were also

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Table 97.3  Acute Toxicity and Plasma Concentrations of Ivermectin and Abamectina Monkey Ivermectin

Humans Abamectin

Therapeutic dose

Ivermectin 0.2 mg/kg

Peak plasma levels

—

—

20 ng/ml

Minimum effect level

2 mg/kg

2 mg/kg

—

Peak plasma levels

110 ng/ml

76 ng/ml

Signs of toxicity

Emesis

Emesis

Toxic effect level

8 mg/kg

8 mg/kg

6.6–8.6 mg/kgb

Peak plasma levels

270 ng/ml

150 ng/ml

Unknown

Signs of toxicity

Emesis

Emesis

Emesis, mydriasis, sedation

Overdose level

24 mg/kg

24 mg/kg

Peak plasma levels

680 ng/ml

390 ng/ml

Signs of toxicity

Emesis

Emesis

Mydriasis

Mydriasis

Sedation

Sedation

a

Lankas and Gordon, 1989. Overdose in humans.

b

Table 97.4  90-Day Dietary Toxicity Studies with Ivermectina, Abamectinb, and Emamectin Benzoatec Study

Exposure (mg/kg/day) Ivermectin

Abamectin

mg/kg SD rat

0.8 d

CD mouse

NS

Beagle dog

2.0

Emamectin

mg/kg

mg/kg

0.4

1.4

0.4

2.5

0.5

NS

8

4

5.4

0.5

0.5

1.0

0.5

0.5

0.25

a

FAO/WHO, 1994. U.S. EPA, 1999a. c U.S. EPA, 1999b. d No study available. b

observed with ivermectin administered to the mrd1a knock­ out mouse (Schinkel et al., 1994, 1995). Furthermore, the incidence of cleft palate correlated with the maternal mortality in a CF-1 mouse study (Lankas et al., 1997). The incidence of cleft palate was also linked to the poly­ morphism of mdr1a in the CF-1 mouse (Umbenhauer et al., 1997).

Developmental toxicity studies have also been con­ ducted with the 8,9-Z-isomer of abamectin in the CF-1 mouse to further evaluate the phenomenon of the linkage of developmental toxicity to the blood–placental barrier (Table 97.8). The NOEL for maternal and developmental toxicity was 0.1 mg/kg/day and 0.05 mg/kg/day, respec­ tively. In young adult CF-1 mice, which were genotyped

Chapter | 97  The Role of P-glycoprotein in Preventing Developmental and Neurotoxicity

2099

Table 97.5  Chronic Dietary Toxicity Studies with Ivermectina, Abamectinb, and Emamectin Benzoatec Study

Exposure (mg/kg/day) Ivermectin mg/kg

Abamectin mg/kg

Emamectin mg/kg

SD rat (105 weeks)d

0.8

0.4

2.0

1.5

2.5

0.25

CD-1 mouse (18 months)

—

—

8

4

5.0 (M)

2.5

7.5 (F) Beagle dog (12 months)

1.0

0.5

0.5

0.25

0.5

0.25

a

FAO/WHO, 1994. U.S. EPA, 1999a. c U.S. EPA, 1999b. b

d

Rat studies with ivermectin and abamectin were only 53 weeks in duration.

Table 97.6  Genotoxicity Studies with Ivermectina, Abamectinb, and Emamectin Benzoatec Mectin

Tests

Mutation

Clastogenicity

Other

Ivermectin

Abamectin

Emamectin

Ames (/ activation)

Negative

Negative

Negative

Mouse lymphoma

Negative

—

—

V-79 Chinese hamster lung (/ activation)

—

Negative

Negative

Mouse bone marrow in vivo

—

Negative

Negative

Chinese hamster ovary in vitro

—

—

Negative

Alkaline elution/rat hepatocyte

—

Negative

Negative

Unscheduled DNA synthesis in human fibroblasts

Negative

—

—

a

FAO/WHO, 1994. U.S. EPA, 1999a. c U.S. EPA, 1999b. b

for their P-glycoprotein expression, the brain concentra­ tions of the isomer 8 h after treatment were 60 times higher in (/) males and females than in the (/) male and female CF-1 mice. Brain concentrations of the delta-8,9isomer of avermectin B1a in (/) CF-1 fetuses were higher than in (/) fetuses, which in turn were higher than in (/) fetuses. To study the development of Pglycoprotein in the placenta in the CF-1 mouse, normal homozygous (/) female and homozygous (/) males were mated and the concentration of P-glycoprotein was measured in the placenta (Lankas et al., 1989). The human population is known to be homozygous positive for this gene. The human fetus is therefore pro­ tected in utero due to an intact placental–blood barrier

with full expression of P-glycoprotein (MacFarland et al., 1994; Nakamura et al., 1997). In addition, this protein has been identified in the capillaries of the brain of the human fetus as early as the third trimester (28 weeks) (Van Kalken et al., 1992) and the level of expression is already the same as that of an adult. Ivermectin has been extensively used worldwide at the high doses administered to humans (in monitored ­ clinical trials as well as in more general therapeutic applications) (FAO/WHO, 1991). No subpopulation of humans demon­ strating variable expression of P-glycoprotein has been iden­ tified. Because humans and other primates have not been shown to have subpopulations deficient in P-glycoprotein, the CF-1 mouse developmental toxicity data are not

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Table 97.7  Developmental Toxicity Studies with Ivermectina, Abamectinb, and Emamectin Benzoatec Study

Dose (mg/kg/day) Ivermectin

Abamectin

Emamectin

LOEL

NOEL

LOEL

NOEL

LOEL

NOEL

CF-1 moused,e

0.2

0.1

0.075

0.05

—

—

Rabbit

6.0

3.0

2.0

1.0

6.0

3.0

SD rat

10.0

5.0

2.0

1.6

4.0

2.0

CF-1 moused

—

0.8c

0.4

0.2

—

—

Rabbit

3.0

1.5

2.0

1.0

—

6.0c

SD rat

—

10.0c

2.0

1.6

8.0d

4.0

CF-1 moused

0.4

0.2

0.4

0.2

—

—

Rabbit

3.0

1.5

2.0

1.0

—

6.0f

SD rat

10.0

5.0

2.0d

1.6

8.0g

4.0

Maternotoxicity

Fetotoxicity

Developmental

a

FAO/WHO, 1994. U.S. EPA, 1999a. c U.S. EPA, 1999b. d CF-1 animals tested were polymorphic for P-glycoprotein (Umbenhauer et al., 1997). e Not evaluated with emamectin benzoate. f No adverse effects at the highest dose tested. g No cleft palates seen at the highest dose tested. b

considered to be of particular relevance for use in human risk assessments for the avermectins (FAO/WHO, 1997). The critical levels and effects for multigeneration reproduction studies with the avermectins are shown in Table 97.9. The LOEL and NOEL values for ivermectin and abamectin were quite similar; the LOEL and NOEL for emamectin benzoate were somewhat higher. Early post­ partum rat pup mortality has been observed with all three avermectins (U.S. EPA, 1999a,b; Lankas et al., 1989). Lankas et al. (1989) observed a significant increase in mortality between days 7 and 14 postpartum in treated dams nursing treated and control pups. In contrast, the mortality, growth, and development of treated and control pups nursing from control dams were similar. Because tox­ icity was only observed in control and treated pups crossfostered to treated dams, it was concluded that neonatal toxicity of ivermectin in rats was a function of postnatal lactation exposure only and not due to in utero exposure. Further, these investigators administered purified, tri­ tium-labeled avermectin B1a (ivermectin B1a) and sampled plasma and milk from dams treated orally with 2.5 mg/kg/ day of radiolabeled ivermectin B1a for 61 days. The pattern

for pup mortality, milk concentration, pup liver, plasma, and brain concentration of ivermectin, and the percentage of adult levels for P-glycoprotein in the pup blood–brain barrier are presented in Table 97.10. Rats also differ from humans by having an increased utilization of fats at the time of parturition (Amano, 1967; Chiu et al., 1986; Scow et al., 1964). This results in a greater release of lipophilic compounds such as abamec­ tin and ivermectin from body fat into milk. In addition, rat milk has a much higher fat content than human milk and this leads to an increased transfer of lipophilic xenobiotic compounds to the neonatal animal compared with what would be anticipated in nursing humans. Ogbuokiri et al. (1993) reported that after a single oral dose of up to 12 mg ivermectin administered to lactating women who were not breastfeeding, the peak concentration in plasma was seen 4 h after treatment. The peak concentration in milk occurred at the same time point, but was two to three times lower than what was seen in plasma. These findings contrast with those found in rats where the concentration of ivermectin observed in rat milk was about three times higher than that in plasma. Based on the uniqueness of the time profile of

Chapter | 97  The Role of P-glycoprotein in Preventing Developmental and Neurotoxicity

2101

Table 97.8  Genotyping Study of CF-1 Mice Treated with 8,9-isomera,b 5 mg/kg

Control

Parameters / F; / M

/ F; / M

/ F; / M

/ F; / M

/ F; / M

Number of fetuses examined

108

105

141

125

127

Number of fetuses with cleft palate

1

0

0

18

80

Number of litters examined

8

9

12

12

12

Number of litters with cleft palate

1

0

0

6

11

Pups with / genotype

19

50

NFc

NF

31

Pups with / genotype

15

NF

NF

41

29

Pups with / genotype

32

NF

39

31

NF

/ With cleft palate

0%

0%

NF

NF

97%

/ With cleft palate

0%

NF

NF

39%

45%

/ With cleft palate

0%

NF

0%

0%

NF

a

Wise et al., 1997. Genotype for P-glycoprotein; , functional gene; , defective gene. c NF, genotype not found. b

Table 97.9  Multigeneration Reproductive Toxicity Studies in SD Rats with the Avermectinsa Ivermectin

Abamectin

Emamectin

LOEL

NOEL

LOEL

NOEL

LOEL

NOEL

Mg/kg/day

0.4

0.2

0.4

0.12

3.6

0.6

Effect

Neonatal mortality

Neonatal mortality

Clinical signs

a

EPA, 1999a,b; Lankas and Gordon, 1989.

P-glycoprotein development (BBB and placenta), and the milk concentration of lipophilic toxicants in rats, it is con­ sidered that it can be safely concluded that these pup deaths are not relevant in terms of human risk assessment.

97.6  Humans: experience with ivermectin Ivermectin has been used clinically for over a decade for the control of Onchocerca volvulus and other parasites in veterinary and human medicine (FAO/WHO, 1993). Onchocerciasis is endemic in large areas of Africa and Latin America. It is estimated that nearly 20 million people are

infected and another 85 million are at risk. Large and diverse populations have been treated with ivermectin in Africa (Burnham, 1993; Chijioke and Okonkwo, 1992; Chippaux et al., 1993; De Sole et al., 1989a,b; Doumbo et al., 1992; Gardon et al., 1997; Ogunba and Gemade, 1992; Pacque et al., 1990, 1991; Whitworth et al., 1991), Polynesia (Cartel et al., 1992), and Latin America (Collins et al., 1992). Typical human doses range from 0.1 to 0.2 mg/kg (FAO/WHO, 1993). The major effect noted following the administration of ivermectin is a severe inflamma­ tory response, called the Mazzotti reaction. The Mazzotti ­reaction is secondary to the efficacy of ivermectin in kill­ ing the microfiliae, which dislodge from their site of ­infestation and are subsequently transported in the blood

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Table 97.10  Pup Mortality, Milk and Pup Tissue Toxicokinetics, and Brain P-Glycoprotein Level Following a Daily Dose of 2.5 mg/kg/day of Tritiated Ivermectin for 61 Daysa Day dose (mg/kg)

Pup mortality

Maternal milk level (g/g)

Pup plasma level (g/g)

Pup liver level (g/g)

Pup brain level (g/g)

Day 1

–0.22/dayb

—

0.094

1.640

0.100

Day 2

6.5

Day 4 Day 5

Rat pup brain p-gp level (%)c

2.324

0.276

3.918

0.251

19.3/day

Day 6

5.7 1.482

0.804

6.106

0.318

Day 7 Day 8

16.9/day

Day 10

4.4 1.052

0.893

6.648

0.264

Day 11

6.9

Day 14

19

Day 15

0.86/day

Day 17

37

Day 20

89

a

Lankas et al., 1989. bCorrected for background by subtracting the pup mortality observed in the control group. c Cukierski (1995); P-glycoprotein expressed as percent of adult level. b

and body fluids (Ackerman et al., 1990). This acute exacerbated immune response can be characterized by pruri­ tus, erythema, edema, vesicle formation, papule formation, and scaling. Adenitis, fever, and hypotension may occur, and severe inflammatory changes may be noted in both the anterior and the posterior segments of the eye. The World Health Organization reviewed reports on the response to treatment for over 26,000 patients admin­ istered ivermectin for parasite control (FAO/WHO, 1993). Single oral doses up to 0.2 mg/kg (bw) produced no major effects except for those resulting from the eradication of the parasite infestation (the Mazzotti reaction). The effects observed in over 200,000 patients treated with ivermectin are summarized in Table 97.11. Although the primary effect noted following the admin­ istration of ivermectin was the Mazzotti reaction, there were two cases of serious neurological response in two patients in a population of 17,877 treated with ivermectin (Gardon et al., 1997). Headache was a common side effect noted, but no association between headache and treatment was observed in a double-blind study on 7148 people con­ ducted by Burnham (Burnham, 1993). During the first year of treatment, pain, edema, itching, and rash were found

s­ tatistically associated with treatment. These reactions diminished in the second year and disappeared by the third year. Hence, in this large human study, patients treated with ivermectin did not exhibit any of the expected neuro­ logical side effects that would have occurred if the blood– brain barrier had been compromised. The populations treated have included not only adults, but also children of all ages and, inadvertently, pregnant women. Epidemiological follow-up of more than 1000 pregnant women treated with ivermectin (primarily in the first trimester) did not yield any indication of an increase in the incidence of miscarriage, stillbirths, or congenital malformations (Burnham, 1993; Chippaux et al., 1993; Pacque et al., 1990). Based on this extensive human data­ base, there should be little concern that neurotoxicity or birth defects might occur in humans exposed to the aver­ mectins at doses less than 0.2 mg/kg.

97.7  Risk characterization Previous joint meetings in 1992 (FAO/WHO, 1992) and 1994 (FAO/WHO, 1994) had established the acceptable

Chapter | 97  The Role of P-glycoprotein in Preventing Developmental and Neurotoxicity

2103

Table 97.11  Observations from Patients Treated with Ivermectin Population treated

Incidence of

Dose (g/kg)

Main effects observed

Reference

14,911

52 (0.35%)

130–200

37 (0.25%) Cases of severe symptomatic postural hypotension

De Sole et al. (1989a)

13 (0.09%) Cases of severe fever 2 (0.01%) Cases of dyspnea 118,925

835 (0.7%)

150

230 (0.19%) Cases of headache

Ogunba and Gemade (1992)

210 (0.17%) Cases of general pains 150 (0.12%) Cases of pruritus 120 (0.10%) Cases of edema 80 (0.06%) Cases of fever 20 (0.02%) Cases of dizziness 15 (0.01%) Cases of vomiting 10 (0.01%) Cases of diarrhea 7,566

992 (13.1%)

150–200

Primarily Mazzotti reaction

Ogunba and Gemade (1992)

460 Cases of headache 50,929

93 (1.83%)

150–200

49 Cases of severe symptomatic postural hypotension

Chijioke and Okonkwo (1992)

34 Cases of severe fever 3 Cases of severe dyspnea 3 Cases of severe pain 7,699

100 (1.3%)

150

daily intake (ADI) for abamectin at 0.0002 mg/kg bw using the NOEL (rat pup toxicity) of 0.12 mg/kg/day derived from the multigeneration reproduction study conducted in Sprague-Dawley rats to which a 500-fold uncertainty factor was applied (Table 97.12). This 500 factor was based on the standard 10 for interspecies differences and 10 for interindividual differences plus an extra 5 due to concern over the teratogenicity of the abamectin delta8,9-isomer in the CF-1 strain of mouse. In 1997, the World Health Organization’s JMPR re-examined the basis for setting the ADI for abamectin and declared that the CF-1 mouse was not suitable for human risk assessment because of its heterozygous (/) or homozygous (/) genetics for P-glycoprotein (FAO/WHO, 1997; Lankas et al., 1997; Umbenhauer et al., 1997). This same rationale, that is, the absence of P-glycoprotein blood–brain barrier in rat pups, postnatally (Betz and Goldstein, 1981; Lankas et al., 1989;

Primarily Mazzotti reaction

De Sole et al. (1989b)

Terao et al., 1996), led the JMPR to readjust the ADI for abamectin to 0.002 mg/day (bw) by reducing the uncer­ tainty factor from 500-to 50-fold for the ADI from the rat multigeneration study. The EPA also declared that the CF-1 mouse was unsuit­ able for human risk assessment but failed to consider the compromised unprotected postpartum period unique to rat pups and a 100-fold uncertainty factor was applied to the NOEL for the multigeneration reproduction study conducted with abamectin (U.S. EPA, 1999a). Temporary tolerances were also set for emamectin benzoate in 1999 (U.S. EPA, 1999b). Unfortunately, the EPA apparently did not apply the same criteria with regard to the inappropri­ ateness of the CF-1 mouse, polymorphic to P-glycoprotein, for human risk assessment. Not only was the NOEL for a 15-day neurotoxicity study in the CF-1 mouse used, but a 300-fold uncertainty factor was also applied to the NOEL

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2104

Table 97.12  Chronic ADI/RfD Values Established for the Avermectins Avermectin

Organization

Study/incidence used

Uncertainty factor

ADI or RfD mg/kg/ day

Ivermectin

JECFA (FAO/WHO, 1992)

Developmental toxicity (CF-1 mouse)

10 intraspecies 10 interspecies

0.001

Abamectin

JMPR (FAO/WHO, 1994)

Multigeneration reproduction (SD rat)

10 Intraspecies 10 interspecies 5 teratogenic concerns

0.0002

JMPR (FAO/WHO, 1997)

Multigeneration reproduction (SD rat)

10 Intraspecies 5—interspecies

0.002

U.S. EPA, 1999a

Multigeneration reproduction (SD rat)

10 Intraspecies 10 interspecies

0.0012

U.S. EPA, 1999b

15-day neurotoxicity study (CF-1 mouse)

10 Intraspecies 10 interspecies 3 short duration of study used

0.00025

Emamectin benzoate

derived from this study in order to calculate the reference dose (RfD). Under appropriate testing conditions, these avermectins are not developmental toxins or reproductive toxins; neither are they genotoxic or carcinogenic. Further, the hazard pro­ files for the three avermectins evaluated are qualitatively and often quantitatively similar. Because of these facts, it would appear to be appropriate to use critical values from the avermectin used extensively clinically for the risk char­ acterization of this class of chemical. Therefore, for acute risk characterization, a clinical dose of 0.2 mg/kg could be used as the NOEL (FAO/WHO, 1993). Conservatively, a 10  interindividual uncertainty factor could be applied, as well as the 3  uncertainty factor proposed by the U.S. EPA, that is, 30  , for an acute RfD of 0.0067 mg/kg. Likewise, using the common mechanism approach for the avermectins, the chronic RfD based on the NOEL for the 1-year dog study (the most sensitive species in the chronic studies) would be suitable. The NOEL for the dog with ivermectin and abamectin (Lankas and Gordon, 1989) and emamectin benzoate (EPA, 1999b), that is, 0.5, 0.25, and 0.25 mg/kg/day, respectively, should be used instead of relying on the data from studies of shorter duration. The chronic RfD of emamectin, based on either study, would be 0.25 mg/kg/day divided by 100 and 3 , or 0.00083 mg/ kg/day, which is essentially the same as the JMPR ADI for abamectin (FAO/WHO, 1997).

97.8  Importance of the   p-glycoprotein blood–brain barrier In the human, the microvasculature, which includes an estimated 40 billion capillaries, serves as a site of chemical

exchange between tissue and blood (Aigner et al., 1997). These capillaries consist of a single layer of endothelial cells (continuous, discontinuous, and fenestrated) sup­ ported by a basement membrane. Capillaries with con­ tinuous endothelial cells are only present in a few organs, including the intestine, bile duct, liver, pancreas, kidney, adrenal, testes, placenta, and brain (Thiebaut et al., 1987). Sites within the brain and the placenta capable of metabol­ ism (Audus et al., 2002; MacFarland et al., 1996; St-Pierre et al., 2004) and active efflux transport by transmembrane pumps such as P-glycoprotein (Fisher and Sikic, 1995; Gottesman and Pastan, 1993). P-glycoproteins are efflux transporters which belong to ATP-binding cassette transporters, the ABC-transporter superfamily (Jones and George, 2004). These efflux trans­ porters are large-membrane proteins (150–180 kDa) consist­ ing of two identical subunits, each with a single adenosine 5-triphosphate (ATP)-binding site and several transmem­ brane domains (Juliano and Ling, 1976). They are highly expressed in endothelial cells at areas that have a barrier function (e.g., the blood–brain barrier and the blood–placental barrier). P-glycoprotein barriers have also been identified in the adrenal gland, colon, testes, and the gravid uterus (Tiirikainen and Krusius, 1991). In addition, its localization along the apical surfaces of the intestines, proximal tubule of the kidney, and the bile ducts of the liver imply that Pglycoprotein is probably involved in secretory functions in humans. P-glycoprotein is a member of a highly conserved mul­ tigene family with isoforms identified in a wide variety of mammalian species, including humans, rats, mice, ham­ sters, pigs, guinea pigs, rhesus monkeys, orangutans, cows, and chickens (Saunders, 1977). Further, P-glycoprotein has been identified in fruit flies (Wu et al., 1991) and in tobacco hornworms and budworms (Lanning et al., 1996).

Chapter | 97  The Role of P-glycoprotein in Preventing Developmental and Neurotoxicity

In addition, P-glycoprotein has been found in mussels and sponges (Kurelec and Pivcevic, 1991), in yeast (McGrath and Varshavsky, 1989), in parasites (Descoteaux et al., 1992), and in nematodes (Lincke et al., 1992). The substrates for P-glycoprotein transporters include large lipophilic molecules that contain at least one aro­ matic ring and a positively charged nitrogen atom. Initially, P-glycoprotein interactions were thought to be limited to only the natural products such as anthracyclines, vinca alkaloids, actinomycin D, epipodophyllotoxins, taxol, and taxotere. However, it is now known that P-glycoprotein also transports steroid hormones, peptide antibiotics, immunosuppressive agents, and calcium channel block­ ers (Ueda et al., 1997) and pesticides including abamectin (Didier and Loor, 1996), ivermectin (Schinkel et al., 1995), 2-acetylaminofluorene and pentachlorophenol (Toomey and Epel, 1995), thiodicarb, and chlorpyrifos (Lanning et al., 1996). P-glycoprotein is encoded as three isoforms. Mouse P-glycoproteins are known as mdr1a (also called mdr3 or Pgy1), mdr1b (also called mdr1 or Pgy2), and mdr2 (Pgy3) (Borst et al., 1993). The isoform mdr1a is primarily found in the intestinal brush border epithelium, the microvessel endothelial cells in the brain and testis, and the microvil­ lus border of the trophoblast and Hofbauer cells of the pla­ centa (MacFarland et al., 1994; Nakamura et al., 1997). The isoform mdr1b is found in the adrenal gland, kidney, and gravid uterus. Both mdr1a and mdr1b are able to pump xenobiotics out of the cells, aiding in multidrug resistance. The isoforms mdr1a and mdr1b of the mouse are compara­ ble to the human MDR1 gene in this regard (Mauad et al., 1994). The isoform mdr2 has been shown to be necessary for bile production and is probably the phosphatidyl cho­ line transporter. Mouse mdr2 is comparable to the human MDR3 (also called MDR2) gene and does not confer multi­ drug resistance to cells. P-glycoprotein transport processes have been conserved across species, probably because this transporter system is essential for adaptation and survival (Saunders, 1977). A population genotypically recessive for P-glycoprotein has not been identified in any species. However, it has been possible to develop a “knockout” mouse for the mdr1a gene (Schinkel et al., 1994, 1995). In addition, polymor­ phism for mdr1a P-glycoprotein gene expression has been reported for the CF-1 mouse (Umbenhauer et al., 1997) and the collie dog (Lankas et al., 1997). Van Kalken et al. (1992) have shown that mdrl/P-glycoprotein is present in human fetal tissues after 7 months through gestation and is born with a full complement of p-glycoprotein. Animals with a recessive mdr1a (/) genotype do not express in P-glycoprotein in endothelial cells at the blood– brain and blood–placental barriers. Studies using knockout mice for mdrla (/) (Schinkel et al., 1994, 1995) have shown that P-glycoprotein is a major determinant of drug efflux transport from the brain. Studies with CF-1 mice

2105

polymorphic for P-glycoprotein (/) (Umbenhauer et al., 1997) lead to a similar conclusion. Mice with (/) P-glycoprotein and CF-1 mice with­ out P-glycoprotein (/) are significantly more sus­ ceptible to the effects of neurotoxicants. Brain levels of ivermectin in the knockout mice that do not express Pglycoprotein were elevated approximately 90-fold more than the wild type [mdr1a (/) genotype] and sensitivity to ivermectin was increased (Schinkel et al., 1994, 1995). Further, CF-1 mice show a unique developmental response to avermectins due to the polymorphic nature of the P-glycoprotein gene. Mdr1a (/) genotype was present in a population of CF-1 mice in a range from 0 to 100%, depending on the genotype of the parental ani­ mals (Umbenhauer et al., 1997). Therefore, experiments carried out with P-glycoprotein substrates in the heteroge­ neous population of the CF-1 mouse must be interpreted with caution and may be unsuitable for risk assessment. Besides the CF-1 mouse model, there are other unique features noted in the standard hazard testing models. The Sprague-Dawley rat pup does not establish a complete P-glycoprotein blood–brain barrier until appropriately 3 weeks postpartum, making it highly vulnerable to neurotoxic effects (Lankas et al., 1989). This is unlike the human fetus, where the blood–brain barrier is formed early on during prenatal development (Jette et al., 1995; Van Kalken et al., 1992; Saunders, 1977) The placenta acts as the barrier between the mater­ nal and fetal circulatory systems and plays an important role in the development and growth of the fetus (Soares, 1987; Knipp et al., 1999). It provides a mechanism for bidirectional exchange of chemicals into and away from the fetus which is critical for normal in utero develop­ ment (Aleksunes et al., 2008). Nutrients and endogenous chemicals are transferred from the maternal circulation to the fetus, thereby providing building blocks for organogen­ esis. Likewise it can restrict fetal exposure to xenobiotic through the transporter systems. The rat placenta is composed of two distinct zones, the junctional zone (invasive and endocrine function) and the labyrinth zone (transport barrier) (Knipp et al., 1999). The junctional zone is adjacent to the maternal compart­ ment and is mainly involved in uterine wall invasion and the production of hormones/cytokines. The labyrinth zone is the main barrier to diffusion and acts to regulate the trans­ fer of nutrients and wastes between the maternal and fetal compartment. The human placenta is villous and the rodent labyrinthine: both are morphologically similar, with mater­ nal blood bathing the trophoblast (see Figure 97.5). Both are morphologically and histologically similar, being both of the hemochorial type, and the syncytial trophoblast cells serve as the transport barriers in both species (Carter, 2001). In addition, fetal membranes consisting of an inverted yolk sac and amniotic membrane extend from the placenta and enclose the fetus. The protective function of the placenta

2106

Figure 97.5  (a) Hemomonochorial villous placenta and (b) rat hemotri­ chorial labyrinthine placenta.

against xenobiotics, which may affect fetal development, is thought to be mediated, at least in part, by transport proteins of the ATP binding cassette (ABC) superfamily, which atten­ uate maternofetal transfer of potentially toxic compounds. Several different transport proteins have been found in placental tissues (Aleksunes et al., 2008; Mitra, 2008; Young et al., 2003). These include proteins already known to be involved in multidrug resistance of tumor cells, like the MDR1 gene product P-glycoprotein (P-gp, ABCB1), the breast cancer resistance protein (BCRP, ABCG2), and several members of the MRP (ABCC) subfamily (MacFarland et al., 1994; Maliepaard et al., 2001; Mao, 2008; Novotna et al., 2004; Pascolo et al., 2003; St-Pierre et al., 2000) as well as organic anion transporting poly­ peptides [OATPS] (St-Pierre et al., 2004). P-glycoprotein actively effluxes drugs from the fetal compartment, dimin­ ishing fetal drug concentrations and preventing drugs from reaching the fetus (Unadkat et al., 2004). The presence or absence of P-gp has been demonstrated to impact upon drug pharmacokinetics (Lin, 2003; Schinkel and Jonker, 2003). For example, the administration during pregnancy of paclitaxel to mdr1a//1b/ knockout mice resulted in up to a 16-fold increased accumulation in the fetus com­ pared with wild-type mice (Smit et al., 1999). In the rodent placenta (Novotna et al., 2004), p-glycoprotein is not detected until after gestation day

Hayes’ Handbook of Pesticide Toxicology

(GD) 12. Figure 97.6 (Western blots) shows that the full complement of P-glycoprotein is not present until GD 17, which is unlike human placentation, where P-glycoprotein has been demonstrated to be maximally expressed in the first trimester during the critical organogenesis phase (Sun et al., 2006; Thiebaut et al., 1987). Consequently there is increased potential for the rodent fetus to be exposed to developmental toxicant during development, in contrast to human fetuses, which develop earlier placental P-glyco­ protein protection. Early chemical exposure (on GD 8 in mice and on GD 9–11 in rats) interferes with neural crest development, leading to palatal shelf hypoplasia or cleft palate (Sulik et al., 1988), suggesting the absence of P-glycoprotein at the time is important in rodents. Fortunately, the human female placenta expresses P-glycoprotein during this crit­ical period of development (Sun et al., 2006; Thiebaut et al., 1987), thereby reducing the potential for the occurrence of such developmental anomalies. Sun et al. found that p-glycoprotein mRNA and pro­ tein levels in tissues obtained between 7 and 13 weeks and between 24 and 35 weeks of gestation were significantly higher than those found at term. The Western blots for P-glycoprotein found in human placenta ratioed against glyceraldehyhe-3-phosphate dehydrogenase used as an internal control for the PCR amplification product normal­ ization are shown in Figure 97.7. Unlike rats and mice, the human blood–brain (Bohr and Mollgard, 1974; Jette et al., 1995; Saunders, 1977; Van Kalken et al., 1992) and blood–placental (MacFarland et al., 1994; Nakamura et al., 1997) P-glycoprotein barri­ ers are formed early in prenatal development and thus the enhanced uptake of abamectin, which has been reported in rodents to be unlikely to occur during the sensitive peri­ ods of neonatal human fetal development. In fact, clinical ­studies based on ivermectin uses in human health show that toxicologically significant uptake of avermectin B1 is not observed in humans. The mectins are selectively toxic to rat, rabbit, and mice fetuses, but only at or near maternotoxic dose levels. This toxicity may manifest itself as developmental abnormality, particularly cleft palate, when there is no P-glycoprotein present in the placenta (Sulik et al., 1988). When the CD-1 mouse fetus is exposed to avermectin B1 during gesta­ tion, during this critical window, before P-glycoprotein is expressed as a barrier protein, or in the absence of Pglycoprotein, (knockout mice, Lankas et al., 1998), the potential for developmental anomalies is increased (Degitz et al., 1998). Mice are the most sensitive species to the effect of ivermectin with maternotoxicity at a dose of 0.2 mg/kg/day (Lankas and Gordon, 1989). These developmentally toxic effects are likely due to enhanced systemic exposure to the fetus arising from lack of P-glycoprotein (an efflux transporter) in the placenta during this early critical phase of organogenesis (Novotna et al.,

Chapter | 97  The Role of P-glycoprotein in Preventing Developmental and Neurotoxicity

2107

Figure 97.6  Percent expression of P-glycoprotein in rodent placenta between GD 11 and 22 (after Novotna et al., 2004).

1.2

P-gp protein (relative expression)

1

MDR–1 Protein in Human Placenta

Conclusion

*

The assessment of the toxicity of the avermectins in the presence or absence of an intact placental–fetal barrier in rodents is an illustrative case study of the effect of species differences in the expression of efflux transporters. Thus, CF-1 mice are more sensitive to the avermectins due to the heterozygous expression of P-glycoprotein. Secondly, the increased postnatal pup mortality in rats is as a result of a lack of expression of P-glycoprotein at birth (BBB and placenta), in contrast to humans, and a high milk con­ centration of lipophilic toxicants, in contrast to humans, and unique to this mouse. The clinical use of ivermectin as well as the additional studies that have been described, provide reassurance that humans are homozygous positive (unimodal) for the P-glycoprotein gene. Consequently, as long as animal models continue to be used to characterize potential risks to the human population, it will continue to be of critical importance to appreciate that the genetic differences underlying the manifestations of toxicity in animal models may not necessarily be appropriate for extrapolation of risk to man.

**

0.8 0.6 0.4 0.2 0 6–13 wks

24–35 wks

Term C.S 38–41 wks

Term Labor 38–41 wks

MDR Figure 97.7  Expression of P-glycoprotein in human placenta from early gestation throughout gestation (after Sun et al., 2006).

2004). In addition, Lankas et al. (1998) showed that fetuses of CF-1 mice lacking the mdr1a gene isoform of P-gp were also susceptible to cleft palate malformation induced by avermectin B1a (Lankas et al., 1998), while homozygous P-glycoprotein dominant wild type were protected from the malformation, further demonstrating the importance of P-glycoprotein as an efflux transporter. In humans and pri­ mates, the first trimester is critical for normal organogenesis. During this period, P-glycoprotein expression is maximal and consequently the fetus is protected. (MacFarland et al., 1994; Nakamura et al., 1997; Sun et al., 2006).

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using a restriction fragment length polymorphism. Toxicol. Appl. Pharmacol. 146, 88–94. Unadkat, J. D., Dahlin, A., and Vijay, S. (2004). Placental drug transport­ ers. Current Drug Metabol. 5, 125–131. U.S. Environmental Protection Agency (EPA) (1999a). Avermectin; pes­ ticide tolerance for emergency exemptions: Final rule. Fed. Reg. 64, 16843–16850. U.S. Environmental Protection Agency (EPA) (1999b). Emamectin ben­ zoate: pesticide tolerance: Final rule. Fed. Rex. 64, 27192–27200. Van Kalken, C. K., Giaccone, G., van der Valk, P., Kuiper, C. M., Hadisaputro, M. M. N., Bosma, S. A. A., Scheper, R. J., Meijer, C. J. L. M., and Pineda, H. M. (1992). Multidrug resistance gene (P-glycoprotein) expression in the human fetus. Am. J. Pathol. 141, 963–1072. Whitworth, J. A. G., Morgan, D., Maude, G. H., Downham, M. D., and Taylor, D. W. (1991). A community trial of ivermectin for onchocer­ ciasis in Sierra Leone: Adverse reactions after the first live treatments rounds. Trans. R. Soc. Trap. Med. Hyg. 85, 501–505. Williams, M., and Yarbrough, G. G. (1979). Enhancement of the in vitro binding and some of the pharmacological properties of diazepam by a novel antihelmintic agent, avermectin B1b. Eur. J. Pharmacol. 56, 1273–1276. Wise, L. D., Lankas, G. R., Umbenhauer, D. R., Pippert, T. R., and Cartwright, M. E. (1997). CF-1 mouse sensitivity to abamectininduced cleft palate correlates with fetal/placental P-glycoprotein genotype. Teratology 55, 41. Wu, C. T., Budding, M., Griffin, M. S., and Croop, J. M. (1991). Isolation and characterization of Drosophila multidrug resistance gene homo­ logs. Mol. Cell. Biol. 11, 3940–3948. Yavanhxay, S. J., Stevens, J. T., Eldridge J. C., Christian, M. S., and Hoberman, A. M. (2004). Ontogeny of P-glycoprotein (PGP) in the brain and gonads of neonatal male and female sprague-dawley rats. The Toxicologist Abstract 1464. Yavanhxay, S. J., Stevens, J. T., Eldridge J. C., Christian, M. S., and Hoberman, A. M. (2005). Ontogeny of P-glycoprotein (PGP) in selected tissues neonatal male and female CD-1 mice. The Toxicologist Abstract 1738. Young, A. M., Allen, C. E., and Audus, K. L. (2003). Efflux transporters of the human placenta. Adv. Drug Deliver. Rev. 55, 125–132.

Chapter 98

DEET Daniel L. Sudakin and Thomas Osimitz Department of Environmental and Molecular Toxicology, Oregon State University, Corvallis, Oregon

98.1  Introduction DEET is an insect repellent that is used in products to prevent bites from insects such as mosquitoes, biting flies, fleas, and small flying insects. DEET was developed by the U.S. Army in 1946 for protecting soldiers in insect-infested areas. Since 1957, insect repellents containing DEET have been widely used by the general public in the United States. It has previously been estimated that approximately 30% of the population uses one or more repellent products containing DEET on an annual basis (U.S. EPA, 1998). The emergence of West Nile Virus in North America may have had an impact on the prevalence of use of DEET in the general population, although it is difficult to obtain current usage statistics. One epidemiological report suggested that communities where West Nile Virus had high prevalence were less likely to have reported using DEET as insect repellent (Gujral et al., 2007). In the United States, DEET is available in several different types of repellent product formulations including formulated lotions, liquids (pump sprays), aerosols, and impregnated materials (such as wrist bands). There are some formulations that combine DEET with sunscreen. A repellent soap containing DEET as active ingredient has been marketed in some regions of the world (Rowland et al., 2004). The concentration of DEET in repellent formulations can range from 4 to 100%. Some studies and literature reviews have suggested that DEET has better repellent efficacy in comparison to other active ingredients (Fradin and Day, 2002; Roberts and Reigart, 2004). Several public health authorities have recommended the use of DEET as one of several strategies to prevent West Nile Virus and other arthropod-borne diseases in adults and children (CDC, 2008; CATMAT, 2005). The repellent mechanism of DEET remains a topic of ongoing investigation. Some studies have suggested that DEET works by forming a vapor barrier with an odor and taste that is offensive to insects (Brown and Hebert, 1997). A frequently cited study concluded that insects are attracted to the lactic acid on human skin, and the vapor from DEET interferes with their ability to locate the lactic acid (Dogan et al., 1999). Other studies have challenged this explanation, finding a repellent effect of DEET with only carbon Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

dioxide as the attractant (Hoffmann and Miller, 2003). A more recent study provided behavioral and other evidence supporting that the repellent efficacy is a result of direct detection and avoidance of DEET in the vapor phase by mosquitoes (Syed and Leal, 2008).

98.2  Chemistry The chemical name for DEET is N,N-diethyl-m-toluamide, and its CAS Registry number is 134-62-3. It is a member of the N,N-dialkylamide family of chemicals. The chemical structure of DEET appears in Figure 98.1. The empirical formula of DEET is C12H17NO, and the molecular weight is 191.26 g/mol. DEET is a liquid with a faint, distinct odor and it is almost colorless. The vapor pressure is 5.6  103 mm Hg at 20° Celsius. Henry’s constant is 2.1  10−8 atm-m3/mol. The octanol/water partition coefficient (log Kow) is 2.02. The soil sorption coefficient (Koc) is 300. DEET is practically insoluble in water (1.0 g/l at 25° Celsius). It is very soluble in ethanol and isopropanol, which are common solvents in repellent formulations containing DEET. DEET is a plasticizer and can damage certain vinyl, plastic, rubber, or elastic materials. DEET has not been reported to damage natural fibers such as wool and cotton (DPMIAC, 2002).

98.3  Overview of toxicology studies There is an abundance of toxicology data that have accumulated on DEET over the past 50 years. Many of the early studies do not meet current regulatory standards for the design, conduct, and interpretation of toxicology studies. More recently, the major manufacturers and formulators of DEET under the auspices of the DEET Joint Venture (DJV) and the Chemical Specialties Manufacturers Association (now known as the Consumer Specialty Products Association) developed a safety and toxicology database as part of the U.S. EPA reregistration process for DEET (U.S. EPA, 1998). 2111

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N,N-diethyl-m-toluamide (DEET) CH3 CH2 N CH3

CH3 CH2

O

Figure 98.1  Chemical structure of DEET.

The studies were conducted in accord with Good Laboratory Practices (GLP), where appropriate, and followed current guidelines for design and interpretation. In addition, several important studies have been published on the pharmacokinetics of DEET in humans. Relatively few studies have been published in the scientific literature describing adverse effects and human poisoning incidents involving DEET, and most of these are limited to case reports and case series. Retrospective reviews of Poison Control Center data and other epidemiological studies involving DEET have been published in the open literature. More recently, several studies have been published that explore the metabolism of DEET by human liver microsomes, and potential interactions from combined exposure to other chemicals including pesticides and consumer products such as sunscreens. Interpreting the toxicological literature on DEET can be challenging for a number of reasons. Many of the toxicology studies have been conducted using ingestion pathways in animal models, a factor limiting their external validity to human exposures when DEET is applied to the skin. Some studies have been based upon in vitro experiments on human skin and liver tissue, which can provide interesting information about mechanisms of absorption and metabolism but have important limitations when considered in the context of intact biological systems. In the past several years, there have been several publications describing adverse effects of DEET at levels of exposure that are much lower than thresholds that have been previously reported in the scientific literature. There are also controversies surrounding the methodology and conclusions reached by some investigators who have published papers describing neurotoxic effects in animal models from chronic low-dose exposure to DEET. The discussions in this chapter include studies that were conducted as part of the comprehensive data development program conducted by the DEET Joint Venture (DJV). Other relevant studies from the open literature are discussed and referenced at the end the chapter. All of the studies conducted by the DJV followed applicable U.S. EPA Testing Guidelines, where applicable, and were conducted in accordance with U.S. EPA Good Laboratory Practice Standards (40 CFR Part 160). Even though the principal route of exposure to DEET

in humans is dermal, most of the studies conducted by the DJV were conducted by the oral route of exposure. The oral route of exposure was selected for many of these studies because little or no toxicity can be demonstrated in laboratory animals at the maximum dose level that can be applied by the dermal route of exposure, i.e., 1000 mg/kg/day. Therefore, to satisfy the EPA Testing Guidelines for establishing a maximum tolerated dose, the oral route of exposure was used for studies examining chronic toxicity and oncogenicity, developmental toxicity, neurotoxicity, and reproductive toxicity. Subchronic exposure, the exposure pattern that is most analogous to human use of DEET products, was evaluated by both the oral and dermal routes.

98.3.1  Acute toxicity studies DEET has a low order of acute toxicity by the oral, dermal, and inhalation routes of exposure. EPA guideline GLP studies conducted on typical production grade DEET by one of the major DEET manufacturers show the rat oral LD50 to be 1892 mg/kg, the rat dermal LD50 was 5000 mg/kg, and the rat 4-h LC50 was  2.0 mg/l (Moore, 2000). DEET produced moderate erythema and edema to the skin of albino rabbits following a 4-h occluded exposure. All skin irritation subsided within 7 days. When instilled into the rabbit’s eye, slight corneal opacity and slight to moderate conjunctive irritation in the form of redness, swelling, and discharge were observed. All ocular irritation cleared within 7 days. In a skin sensitization study conducted using the Buehler guinea pig method, no evidence of skin sensitization was observed. Similar findings on the acute toxicity of DEET have also been presented in publications in the open literature (Macko and Weeks, 1980; Weil, 1973). In an acute oral toxicity study conducted in rats of different ages, it was shown that the acute oral LD50 of DEET increased four- to fivefold between the ages of 11 days and 47–56 days of age (Verschoyle et al., 1992). Two publications describe findings in cats and dogs that were simultaneously administered DEET and fenvalerate, a combination of ingredients used in a commercial tick and flea spray product for dogs at the time the studies were conducted (Dorman et al., 1990; Mount et al., 1991). Three other publications describe studies in which potential interactions between DEET, permethrin, and/or pyridostigmine were evaluated (Abou-Donia et al., 1996; Chaney et al., 1999; McCain et al., 1997). The interest in this combination of chemicals arose from concerns about military personnel in the Persian Gulf War that potentially could have been exposed simultaneously to these chemicals. While the findings presented in these publications provide suggestive evidence for possible interactions between DEET and these other chemicals, the evidence for actual chemical synergy is far from conclusive. Also, because DEET was administered to animals orally, subcutaneously or intraperitoneally at or near lethal dose levels, the

Chapter | 98  DEET

external validity of these findings to normal human exposure is limited.

98.3.2  Subchronic toxicity studies A number of subchronic studies were conducted as part of the DJV Data Development Program. These studies were conducted to meet regulatory requirements for 90-day oral and dermal toxicity studies, and for dose range-finding for subsequent longer studies. These studies also addressed findings from other subchronic investigations of DEET in different animal models.

98.3.2.1  Rat 90-Day Dietary Toxicity The Johnson (1987b) study was designed to assess the potential for DEET to produce subchronic oral toxicity and to develop data that could be used to select dose levels for a rat two-generation reproduction study and a rat chronic toxicity/oncogenicity study. DEET was incorporated into the diet and administered to Charles River CD rats at concentrations such that dose levels of 0, 100, 500, 1000, 2000, and 4000 mg/kg/day could be evaluated. Fifteen male and 15 female rats were evaluated at each level. Due to rejection of the treated diets and subsequent body weight depression and mortality, the 4000 mg/kg/day group was discontinued after 7 days. Animals in the remaining groups continued through to the end of the 90-day test period. At the dose levels that were evaluated for 90 days, no treatment-related clinical signs or effects on hematology, clinical chemistry, or gross pathology were observed. Decreased body weight and food consumption were observed at dose levels 500 mg/kg/day and signs of inanition were evident at 2000 mg/kg/day. Nonspecific liver weight increases without correlative microscopic findings occurred in all treatment groups except the 100 mg/kg/day females. Renal lesions were observed microscopically in males at all concentrations and included granular cast formation, multifocal chronic inflammation, regenerative tubular epithelium, and hyaline droplets. These renal lesions were considered to be reflective of alpha2u-globulin induced nephropathy, a condition which is unique to certain strains of male rats and with no relevance to human risk (Goldenthal, 1992).

98.3.2.2  Mouse 90-Day Dietary Toxicity The purpose of the Johnson (1987a) study was to inform the selection of dose levels for an 18-month mouse chronic toxicity and oncogenicity study. DEET was incorporated into the diet and administered to Charles River CD-1 mice at concentrations such that dose levels of 0, 300, 1000, 3000, 6000, and, 10,000 mg/kg/day could be evaluated. Fifteen male and 15 female mice were evaluated at each level. Due to rejection of the treated diets, the 6000 and 10,000 mg/kg/day dose groups were discontinued after 2 weeks. Animals in

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the remaining groups continued through to the end of the 90-day test period. At the dose levels that were evaluated for 90 days, no treatment-related clinical signs or effects on food consumption or gross pathological observations were noted in any dose group. Decreased defecation early in the study and decreased body weight gain were observed in animals in the 3000 mg/kg/day dose group. Absolute and relative liver weights were increased in both males and females at dose levels 1000 mg/kg/day and in females at 300 mg/kg/day. Multifocal hepatocellular hypertrophy was observed at a high incidence in males and females at 3000 mg/kg/day and at a lower incidence in females at 1000 mg/kg/day. The increased liver weights and the corresponding hypertrophy were considered to be adaptive changes rather than an indication of systemic toxicity. No other gross or microscopic changes were observed. Hematology and clinical chemistry evaluations were not included in this dose range-finding study.

98.3.2.3  Hamster 90-Day Dietary Toxicity The Goldenthal (1989b) study was conducted to determine if the renal lesions observed in the rat 90-day dietary toxicity study could be produced in other species. DEET was incorporated into the diet and administered to Golden Syrian VAF/Plus hamsters at 0, 1000, 5000, 10,000, and 15,000 ppm. Fifteen male and 15 female hamsters were evaluated at each dietary concentration. Clinical signs including labored breathing, decreased defecation, decreased activity, pale skin, and mortality were observed at 15,000 ppm. Decreased body weights and food consumption were observed in males at 5000 ppm and in males and females at  10,000 ppm. The testes and epididymides appeared smaller and decreased testis weights were observed at  10,000 ppm. These observations were accompanied by an increased incidence of tubular degeneration in the testes and an associated accumulation of cellular luminal debris in the epididymides. Blood potassium levels were elevated at 15,000 ppm. No other effects on the hematological or clinical chemistry parameters were observed. Most significant, the renal lesions observed in the rat 90-day dietary toxicity study were not observed in this study.

98.3.2.4  Rat 90-Day Dermal Toxicity DEET was applied dermally to the shaven backs of Charles River CD rats 5 days per week for 13 weeks at dosage levels of 0, 100, 300, and 1000 mg/kg/day. The 1000-mg/kg/day dose level represented the maximum dose of DEET that could be applied dermally without significant runoff. Fifteen male and 15 female rats were evaluated at each dosage level. Treatmentrelated effects included dermal irritation, body weight depression in males at the 1000 mg/kg/day dose level, and renal lesions that were observed in males only at all dose levels. Microscopically, these lesions were described as granular

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cast formation, multifocal inflammation, regenerative tubular epithelium, and hyaline droplets. The dermal irritation was confirmed microscopically in the form of acanthosis and/or hyperkeratosis. The renal lesions were accompanied by elevated kidney weights and slightly increased urea nitrogen levels at 13 weeks at the 300 and 1000 mg/kg/day dose levels. Increased liver weights were also observed in male and female rats in the 1000-mg/kg/day dose level. No treatmentrelated clinical signs or effects on food consumption, hematology, or ophthalmology were observed in any of the treatment groups. As was the case in the 90-day oral toxicity study, the renal lesions observed in the male rats were attributed to alpha2u-globulin nephropathy and were not considered relevant for humans (Johnson, 1987c).

98.3.2.5  Micropig 90-Day Dermal Toxicity The purpose of the Goldenthal (1991) study was to evaluate the toxicity of DEET following dermal exposure in a nonrodent species and to develop data to demonstrate that the renal lesions observed in the rat 90-day oral and dermal toxicity studies do not occur in nonrodents. DEET was applied dermally to the shaven backs of Charles River micropigs 5 days per week for 13 weeks at levels of 0, 100, 300, and 1000 mg/kg/day. Four male and four female micropigs were evaluated at each level. Parameters evaluated included observations for clinical signs, dermal irritation, body weight, food consumption (approximated), hematology, clinical chemistry, organ weights, and gross and microscopic pathology. With the exception of slight skin irritation at the application site in all treatment groups, no treatment-related effects were noted in this study. The grossly observed skin irritation was confirmed microscopically in the form of acanthosis and/or hyperkeratosis.

98.3.2.6  Subchronic Studies to Evaluate Further Male Rat Kidney Lesions The findings from the rat 90-day oral and dermal toxi­city studies discussed above indicated that DEET produces kidney lesions in male Charles River CD rats that are characteristic of the renal lesions produced by a wide range of chemicals that induce alpha2u-globulin accumulation in the epithelial cells of renal proximal tubules. These lesions were not observed in female CD rats nor in animals of any other species in which 90-day toxicity studies were conducted. In order to investigate further the correlation between the renal findings observed in the rat subchronic toxicity studies and alpha2u-globulin, two additional 90-day studies were undertaken by the DJV. The first study investigated renal effects of DEET in three different strains of male rats, two strains that produce aplha2u-globulin and one that does not produce alpha2u-globulin (Goldenthal, 1992). The second study investigated the effect of castration on the renal toxicity of DEET to male rats. This study was of interest because the

Hayes’ Handbook of Pesticide Toxicology

synthesis of alpha2u-globulin is influenced by the level of circulating androgens (Goldenthal, 1989a). The results of these and other studies support the conclusion that the renal lesions observed in male rats following DEET administration are associated with accumulation of alpha2u-globulin. The findings of these studies showed that microscopic changes observed are characteristic of the lesions produced by other chemicals that produce these effects via this mechanism. They also demonstrated that the lesions were not observed in DEET-treated female rats or in mice, hamsters, dogs, or micropigs of either sex. It was also shown that castration modifies the response, and similar lesions were not observed in a strain of male rat that does not produce alpha2u-globulin. Because alpha2u-globulin induced nephropathy is not observed in humans, the renal lesions observed in male rats in several DEET studies are not considered to be relevant to human risk assessment.

98.3.3  Developmental toxicity 98.3.3.1  Rat Developmental Toxicity Dose range-finding and definitive developmental toxicity studies were conducted with Charles River CD rats. In both studies, DEET was administered undiluted by oral gavage on gestation days 6–15. Control animals received corn oil at the volume used in the highest dose group in each study. In the dose range-finding study, five mated female rats per group were evaluated at levels of 0, 62.5, 125, 250, 500, and 1000 mg/kg/day. In the definitive study, 25 mated female rats per group were evaluated at levels of 0, 125, 250, and 750 mg/kg/day. Parameters evaluated in both studies included observations for clinical signs, body weights, food consumption, maternal liver and gravid uterine weights, maternal ovarian and uterine exams, and fetal external examinations. In the definitive study, the fetuses also were given detailed internal soft-tissue and skeletal examinations (Schoenig et al., 1994). In the dose range-finding study, maternal toxicity in the form of mortality, decreased body weights, decreased food consumption, hypoactivity, ataxia, prostration, unkempt appearance, urine stains, and perioral wetness was observed at 1000 mg/kg/day. No maternal effects were observed at levels below 1000 mg/kg/day and no evidence of developmental toxicity was observed at any dose. On the basis of these results, dose levels of 0, 125, 250, and 750 mg/kg/day were selected for the definitive study. In the definitive study, maternal toxicity, including mortality, decreased body weight gain, decreased body weights, decreased food consumption, hypoactivity, ataxia, decreased muscle tone, urine stains, foot splay, perinasal encrustation, perioral wetness, and increased liver weights, was observed at 750 mg/kg/day. The only fetotoxic effect at this clearly toxic maternal dose level was a reduction in fetal body weights per litter. The incidence of external, visceral, and skeletal variations and/or malformations were comparable in the control and treatment groups.

Chapter | 98  DEET

98.3.3.2  Rabbit Developmental Toxicity Dose range-finding and definitive developmental toxi­city studies were conducted in female New Zealand White rabbits. In both studies, DEET was administered undiluted by oral gavage on gestation days 6–18. Control animals received corn oil at the volume used in the highest dose group in each study. In the dose range-finding study, five mated female rabbits per group were evaluated at levels of 0, 62.5, 125, 250, 500, and 1000 mg/kg/day. In the definitive study, 16 mated female rabbits per group were evaluated at levels of 0, 30, 100, and 325 mg/kg/day. Parameters evaluated in both studies included observations for clinical signs, body weights, food consumption, maternal liver and gravid uterine weights, maternal ovarian and uterine exams, and fetal external examinations. In the definitive study, the fetuses also were given detailed internal soft-tissue and skeletal examinations (Schoenig et al., 1994). In the dose range-finding study, clinical signs of maternal toxicity in the form of mortality, hypoactivity, ataxia, slow or rapid respiration, and gasping were observed at 1000 mg/kg/day. Mortality and rapid respiration also were noted at 500 mg/kg/day and rapid respiration was observed at 250 mg/kg/day. Decreased body weight gain and food consumption were observed at 500 mg/kg/day. There were no signs of maternal toxicity at 125 mg/kg/day and below and no signs of fetotoxicity or developmental toxi­ city at any dose level. On the basis of these results, dose levels of 0, 30, 100, and 325 mg/kg/day were selected for the definitive study. Maternal toxicity in the form of decreased body weight gain and food consumption was observed in the 325 mg/kg/day group. No evidence of fetotoxicity was observed and the incidence of external, visceral, and skeletal variations and/or malformations were comparable in the control and treatment groups.

98.3.4  Reproductive Toxicity 98.3.4.1  Rat Two-Generation Reproduction Study DEET was incorporated into the diet and administered to the rats at concentrations of 0, 500, 2000, and 5000 ppm. The F0 parental generation consisted of 28 males and 28 females per group, which were administered treated or control diet for at least 80 days prior to mating. Twenty-eight male and 28 female offspring per group from the F1 generation were selected randomly to become the parents of the F2 generation. These animals were treated for at least 93 days prior to mating. For both parental groups, treatment was continued through gestation and lactation. Parameters evaluated in the parental rats included observations for clinical signs, body weight and food consumption measurements, gross necropsy, and microscopic examination of gross lesions and reproductive tract organs. Reproductive

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and litter parameters that were evaluated included male and female fertility indices, events at parturition, gestation duration, litter size, numbers of viable and stillborn pups, and pup survival and growth during lactation (Schardein, 1989). Parental toxicity in the form of decreased body weight and food consumption was noted for males and females in the F0 and F1 generations at 5000 ppm, and decreased body weight was noted for males in the F0 generation at 2000 ppm. A slight increase in hair loss was observed in the F0 and F1 females at 5000 ppm. In the F1 adult males, kidney lesions including mottled kidneys, hyaline droplets, chronic inflammation, regenerative tubules, and renal granular cast formation were observed. These kidney effects occurred in a dose-related manner in all treatment groups and were characteristic of alpha2u-globulin nephropathy. No other treatment-related effects were observed in the parental generations at 500 ppm. Neonatal toxicity as evidenced by reduced pup sizes and weights in both generations was noted for males and females in the 5000 ppm group. No treatmentrelated effects were observed in pups at 2000 ppm. No treatment-related effects on reproduction or fertility were observed at any of the dose levels evaluated in this study.

98.3.4.2  Reproductive Toxicity Studies Referenced in the Open Literature Three other nonguideline, non-GLP studies that address reproductive toxicity are described in the open literature. One study was a subchronic study conducted in rats by the subcutaneous route of exposure that was designed to evaluate dominant lethal effects and male fertility (Wright et al., 1992). A second was a rat subchronic inhalation study in which testicular weights and sperm morphology were evaluated (Macko and Bergman, 1980). The third was a rat subchronic dermal toxicity study in which sperm morphology was evaluated (Brusick, 1980). An increase in the incidence of abnormal sperm was reported in the latter two studies; however, the evidence for this finding is very weak.

98.3.5  Chronic Toxicity and Oncogenicity 98.3.5.1  Dog Chronic Toxicity DEET was administered orally for 1 year to purebred beagle dogs via gelatin capsules at 0, 30, 100, and 400 mg/kg/day. The daily dose of DEET was divided equally into two doses administered in the morning and afternoon following a 1-h period of food availability. Four male and four female dogs were evaluated at each dose level. Parameters evaluated in this study included observations for clinical signs, body weight and food consumption measurements, hematology, clinical chemistry, urinalysis, ophthalmology, organ weights, and gross and microscopic pathology. Treatment-related effects were observed only in the 400-mg/kg/day group and included emesis, ptyalism, decreased body weights, and decreased

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food consumption for both males and females. One male in the 400 mg/kg/day group also exhibited occasional ataxia, tremors, abnormal head movements, and convulsions. These clinical signs generally occurred shortly after dosing and were followed by complete recovery before the succeeding dose. Other treatment-related effects observed in the 400-mg/kg/day group included transient reduction in hemoglobin and hematocrit levels, increased alkaline phosphatase levels (males only), decreased cholesterol levels, and increased potassium levels (males only).

98.3.5.2  Mouse Oncogenicity Study DEET was incorporated into the diet and administered for 78 weeks to Charles River CD-1 mice at concentrations such that dose levels of 250, 500, and 1000 mg/kg/day could be evaluated. Sixty male and 60 female mice were evaluated at each dose level. In addition, two independent untreated control groups, each consisting of 60 male and 60 female mice, were included in the study. Parameters evaluated included observations for clinical signs and palpable masses, measurements of body weight and food consumption, hematology, organ weight measurements, and gross and microscopic pathology. A slight decrease in body weight and food consumption was noted at the 1000 mg/kg/day dose level and an increase in liver weights was noted in male and female mice at 500 and 1000 mg/kg/day. The liver weight increases were considered to be adaptive in nature, since no corroborative microscopic findings were observed. No other treatment-related effects were observed and DEET administration had no effect on survival or tumor incidence (Schoenig et al., 1999).

98.3.5.3  Chronic Toxicity/Oncogenicity Studies Referenced in the Open Literature Two other chronic toxicity/oncogenicity studies are reported in the open literature. In one of these studies, DEET was applied dermally to mice for 120 weeks (Stenbäck, 1976). DEET was also applied dermally to the ears of rabbits for 95 weeks (Stenbäck, 1976). No treatment-related increase in tumors was observed in either study. Both studies had many major deficiencies relative to today’s standards for conducting chronic toxicity/oncogenicity studies.

98.3.6  Neurotoxicity 98.3.6.1  Acute Neurotoxicity Study The neurotoxicity potential of DEET was evaluated in Charles River Crl:CD VAF/Plus rats. A single dose of un­­ diluted DEET was administered by oral gavage at dose levels of 0, 50, 200, or 500 mg/kg and the rats were observed for 14 days following dose administration. Ten male and 10 female rats were evaluated at each dose level. Parameters evaluated included observations for clinical signs and measurements of

Hayes’ Handbook of Pesticide Toxicology

body weight and food consumption. In addition, functional observational battery (FOB), thermal response test, and motor activity measurements were made at 1 h, 24 h, and 14 days following dose administration (Schoenig et al., 1993). No clinical signs of toxicity were observed in any animals and there were no treatment-related effects on body weight or food consumption. No effects related to DEET exposure were observed in the FOB. Animals in the 500 mg/kg dose group exhibited an increased response time to heat in the thermal response test and slightly decreased rearing activity in the motor activity test. These effects occurred in animals of both sexes at the 1-h posttreatment observation interval but were not observed at 24 h or 14 days after treatment. No other treatment-related effects were observed during this study.

98.3.6.2  Chronic Neurotoxicity Study The neurotoxicity potential of DEET was evaluated following multigeneration plus chronic administration. The animals used in this study were second generation (F2) offspring from the rat multigeneration reproductive toxicity study in which DEET was administered continuously over two generations at dietary concentrations of 0, 500, 2000, and 5000 ppm (Schoenig et al., 1993). Following weaning, all control offspring and two male and two female F2 generation pups from 21 to 25 litters per treatment group were selected and maintained on the treated diets for an additional 9 months. During the 9-month dietary administration period, the rats were observed for clinical signs, and body weight and food consumption measurements were made. At the end of this 9-month period, one male and one female from each of 20 litters were selected for the neuro­ toxicity evaluations. An additional control group of the 10 males and 10 females was selected from the rats maintained on basal diet for use as a sham group in the passive avoidance test. Neurotoxicity evaluations included a functional observational battery, motor activity testing, M-water maze, acoustic startle habituation, and passive avoidance. In addition, comprehensive neuropathological examinations were performed on one male and one female from each of 10 litters per group. At the time that these neurotoxicity evaluations were performed, all animals were approximately 40 weeks of age. No clinical signs of toxicity were noted throughout the 9-month dietary administration period; however, decreased body weights relative to the control animals were observed for all treatment groups. These findings were somewhat unexpected at 500 and 2000 ppm, since decreased body weights in the original multigeneration study were observed only at 5000 ppm. The decreased body weights at 500 and 2000 ppm may have been due to random selection of the animals at study start, since the treatment groups were not balanced with respect to animal weight at the time of selection. In the in-life neurotoxicity evaluation, the only treatmentrelated finding was a transient increase in exploratory locomotor activity at 5000 ppm. Equivocal findings were noted in

Chapter | 98  DEET

the M-water maze, in which a decrease in initial choice accuracy on reversal was observed in all of the treated groups as compared with the controls. However, the effect was not dose-dependent, and no confirmatory evidence was found for a DEET-related reversal learning effect on response times and total errors, which are the primary measures of performance recorded for this task. Therefore, this finding was not considered to be adequate evidence of an effect of DEET on learning. No treatment-related effects were observed in the comprehensive neuropathological examinations performed on the central and peripheral nervous tissues. Since the only potential neurotoxic effect that was observed in this study was observed at 5000 ppm, a dose level at which other toxic effects have been observed, the results of this study demonstrate that the nervous system does not appear to be a selective target when DEET is administered chronically at dose levels up to and including the MTD. In addition, the results of this study demonstrate that the chronic administration of DEET at an MTD dose does not result in any morphologic changes in the tissues of the nervous system.

98.3.6.3  Other Studies Describing Neurotoxic Effects in Laboratory Animals Other studies describing neurotoxic effects and neuropathological findings associated with DEET have been published in the open literature. Several of these publications have originated from one laboratory and have described perplexing findings that are inconsistent with the results of other investigators. In one subchronic toxicity study of adult rats receiving a low, daily dermal dose of 40 mg/kg DEET in 70% ethanol for 60 days, extensive neuronal degeneration was observed in the motor cerebral cortex, the dentate gyrus, hippocampal cells, and the Purkinje cell layer of the cerebellum (Abdel-Rahman et al., 2001). Remarkably, the authors reported there were no observable clinical differences between treated animals and controls. This publication did not discuss the lack of clinical correlation with the findings of diffuse neuronal cell death. Another publication by the same investigators reported abnormalities in sensorimotor function in rats after chronic dermal exposure to DEET at doses 10 times lower than their previous investigation (4 mg/kg/day) (Abou-Donia et al., 2001). The authors reported this dose was associated with a significant decrease in beam walk scores in comparison to controls; however, another study by the same authors using a much higher dose (40 mg/kg/day) did not observe any significant abnormalities in beam walk scores (AbdelRahman et al., 2001). In addition to the inconsistent observations in these studies, the doses that were reportedly associated with abnormal beam walk scores (4 mg/kg/day) are lower than daily estimated dermal exposure estimates that have been described for DEET in humans, ranging from 9.68 to 37.63 mg/kg/day (U.S. EPA, 1998). Questions relating to the methodology and conclusions reported by these studies of low-dose dermal exposure to

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DEET have been published in the open literature (Schoenig, 2002). More recently, the validity of the neuropathological findings reported in these studies have also been seriously questioned, as they appear to be consistent with a “dark neuron” effect resulting from postmortem manipulation or trauma in brain tissue (Jortner, 2006). Given the inconsistent findings and the questions that have been raised in relation to the accuracy of the neuropathology results, the external validity of these studies of low-dose dermal exposure to DEET is dubious.

98.3.7  Genotoxicity studies 98.3.7.1  Salmonella/Mammalian-Microsome Plate Incorporation Assay (Ames Test) Five strains of Salmonella typhimurium, TA98, TA100, TA1535, TA1537, and TA1538, were used. Each strain was tested in the presence and absence of metabolic activation by a rat liver S-9 system induced with Aroclor 1254. The concentrations of DEET evaluated in these studies (with and without metabolic activation) ranged from 278 to 8333 g/plate. Mutagenic frequency did not increase in any of the tester strains. Results from the initial assay were confirmed in an independent repeat assay (San and Schadly, 1989).

98.3.7.2  Genotoxicity Studies Referenced in the Open Literature One study evaluated reverse gene mutation in Salmonella typhimurium (Zeiger et al., 1992) and another evaluated reverse gene mutation in both Salmonella typhimurium and Saccharomyces cerevisiae (Brusick, 1976). A dominant lethal assay in Swiss white mice was also investigated in one study (Swentzel, 1978). No significant activity was observed in any of these studies.

98.4  Pharmacokinetic studies: animals and humans 98.4.1  Absorption, Distribution, Metabolism, and Excretion in Rats Radiolabeled [14C]DEET was administered orally and dermally to rats to evaluate absorption, distribution, metabolism, and excretion. Tissues were analyzed for 14C residues 7 days following oral and dermal treatments. When administered orally, the percent of 14C in the circulation and tissues was higher than when applied to the skin. Radiolabeled DEET was consistently detected at higher levels in the liver, kidney, and fat tissues, in comparison to plasma (Schoenig et al., 1996). DEET was completely metabolized after oral and dermal administration (Schoenig et al., 1996). Mass

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spectroscopy identified two major urinary metabolites of DEET, m-[(N,N-diethylamino)carbonyl]benzoic acid and m-[(ethylamino)carbonyl]benzoic acid. When administered orally, within 7 days 85–91% of the administered [14C]DEET was detected in the urine and 3–5% in the feces. The rates of excretion into urine depended on the dosing regimen. [14C]DEET was excreted most rapidly when administered as a repeated oral low dose versus a single oral low dose, suggesting that when administered repeatedly, enzymes that metabolize DEET were induced. The rate of DEET excretion was slowest when administered orally as a single high dose. When radiolabeled DEET was administered dermally at a low dose, it underwent predominantly urinary excretion within 7 days (74–78%), with a relatively lower proportion being excreted in the feces (4–7%), and 6.5% remaining on the skin surface. The rate of excretion was slower when administered on the skin in comparison to oral administration (Schoenig et al., 1996). These findings are consistent with earlier studies of radiolabeled [14C]DEET, which reported that after dermal application to mouse skin some of the radioactivity was retained in the superficial skin, even after washing with ethanol (Blomquist and Thorsell, 1977).

98.4.2  Absorption, Metabolism, and Excretion in Humans Interpreting the literature on DEET absorption is challenging, because the available data provide evidence that skin permeation is significantly influenced by the solvent(s) in which it is formulated. An in vitro study of human skin reported that the total amount of DEET permeated from a solution in 30–45% ethanol was significantly higher than that of a pure DEET solution (Stinecipher and Shah, 1997). The encapsulation of DEET in a liposphere system has been reported to reduce transdermal absorption and increase the protection time in a rabbit model and in vitro studies of human skin (Domb et al., 1995; Iscan et al., 2006; Kasting et al., 2008). Other formulations investigated include a polyethylene glycol and polyacrylic acid polymer system, which have been reported to reduce transdermal absorption and enhance repellence in a canine model, in comparison to an ethanol solution containing the same concentration of DEET (Qiu et al., 1997). A “wash-in” effect, defined as enhancement of percutaneous absorption as a result of aggressive skin contamination or washing, has been reported to occur in some in vitro investigations of DEET in repellent formulations (Moody and Maibach, 2006). When human skin was tested in vitro with three different commercial formulations containing DEET (14–95%), increased permeation was observed in all samples when soap was used to wash the skin 24 h after application (Moody et al., 1995). The “wash-in” effect of DEET has not been evaluated in vivo in experimental studies of humans.

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Relatively few studies have been conducted on the pharmacokinetics of DEET in humans. An early study of dermal absorption in four volunteer subjects reported that 16.7% of a single dose of DEET to the forearm was systemically absorbed (Feldmann and Maibach, 1970). The dermal absorption of DEET in male volunteers was subsequently investigated in another study (Selim et al., 1995). In this investigation, 12–15 mg of radiolabeled DEET was applied to a 4  6-cm area of the forearm. The material was left in contact with the skin for 8 h, and subsequently rinsed from the skin. The concentrations of radiolabeled DEET that were investigated were either technical grade DEET or a 15% (w/w) solution of DEET in ethanol. Dermal absorption was 5.63% of the applied dose for the group receiving technical grade DEET, and 8.41% of the applied dose for subjects receiving 15% DEET in ethanol. Most of the absorbed dose was excreted within 12 h of application, and elimination occurred predominantly in the urine. Excretion of the applied dose was nearly complete at 24 h after application. Osimitz et al. (unpublished) reported on the plasma profile of DEET in humans following single and repeated dermal applications at the 95th percentile of typical human use. In this study, three male and three female human volunteers were administered undiluted DEET by the dermal route of administration at the 95th percentile of human use (3 g/day for females and 4 g/day for males). Both single and repeated (8 h per day for 4 consecutive days) applications were evaluated. DEET plasma levels were profiled on the 1st and 4th days of the study. The DEET plasma profiles for humans under this scenario are summarized in Table 98.1. The findings from this study show that DEET does not appear in the blood of humans until 1–2 h after dermal application, after which time the DEET plasma levels gradually increase until the material is washed off 8 h after application. The DEET plasma profiles and peak plasma levels were similar in males and females and did not increase after repeated dosing for 4 consecutive days. The overall mean peak plasma level was 0.45 g/ml. The overall mean area under the DEET plasma concentration vs. time curve (AUC) was 3.51 g h/ml. DEET metabolites have been characterized in some human studies after dermal administration. In one study, DEET appeared to be completely metabolized as the parent compound was not detectable in urine samples (Selim et al., 1995). As many as six metabolites of DEET were detected using high-performance liquid chromatography (HPLC) analysis, with no apparent differences in metabol­ ite profiles in study subjects receiving different formulations. The study reported less than 0.1% of the applied dose was measured in tape strips applied to superficial skin, suggesting that accumulation of DEET in the epidermis was not of pharmacokinetic significance. Two of the main human urinary metabolites identified in this study were m-[(N,N-dieethylamino)carbonyl]benzoic acid and m-[(ethylamino)carbonyl]benzoic acid.

Chapter | 98  DEET

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Table 98.1  DEET Plasma Levels in Human Volunteers Administered DEET Dermally at the 95th Percentile of Human Use DEET exposure metric

Single exposurea

Repeated exposureb

Male

Female

Mean

Male

Female

Mean

0.62

0.43

0.52

0.49

0.25

0.37

5.11

3.07

4.09

3.76

2.10

2.93

Mean maximum Plasma level (g/ml) c

AUC (g hr/ml) a

From data collected during and following the first of four daily dermal applications of technical grade DEET at dose levels of 4 g/day for males and 3 g/day for females. b From data collected during and following the fourth of four daily dermal applications of technical grade DEET at dose levels of 4 g/day for males and 3 g/day for females. c AUC  area under the DEET plasma concentration versus time curve.

Studies of repeated dermal exposure to DEET in field studies have been infrequently reported. One investigation of repeated dermal application of DEET was conducted in a small sample of National Park Employees who worked in hot, humid environments and were exposed to approximately 1 g of DEET (71% active ingredient) on the skin on a daily basis (Smallwood et al., 1992). A mid-week urine specimen was obtained from eight workers at the end of the work shift. High-performance liquid chromatography of these urine samples revealed detectable levels of DEET in six individuals, with concentrations ranging from 0.58 to 5.69 g/ml. An analysis for DEET metabolites was not reported. In another study involving higher levels of exposure under experimental conditions, a human subject applied 10.4 g of a commercial formulation of DEET to 75% of the skin surface (133 mg/kg), and urinary metabolites were identified over a 36-h period. DEET was detected in the urine for 18 h after exposure, and additional metabolites were identified including products of oxidation of the methyl group in the meta position, and dealkylation products of the amide side chain. These metabolic transformation reaction products are similar to findings from other studies of DEET metabolism (Selim et al., 1995). Based upon metabolic transformation reactions that have been observed in animal and human studies, at least two major metabolic pathways have been identified for DEET (Figure 98.2). One reaction involves oxidative hydroxylation of the aromatic methyl group in the meta-position, yielding N,N-diethyl-m-hydroxymethylbenzamide (BALC). The other major reaction involves dealkylation of an N-ethyl group, producing N-ethyl-m-toluamide (ET). Subsequent oxidation, hydroxylation, and glucuronidation reactions may also occur, which might explain additional metabolites (such as N,N-diethyl-m-carboxylbenzamide) that have been reported in human studies. The specific cytochrome P450 isoforms responsible for DEET metabolism in humans has been the subject of some investigations. One in vitro study of pooled human

hepatic microsomes reported the P450 isoforms having the highest activity toward formation of BALC metabol­ ites were CYP2B6 and CYP1A2 (Usmani et al., 2002). Isoforms CYP2C19, CYP3A4, 3A5, and 2A6 produced detectable amounts of ET, with CYP2C19 having a significantly higher level of activity. A significantly higher affinity (lower Km) was reported for the ring methyl hydroxylation reaction (BALC) in comparison with N-deethylation (ET). Considerable interindividual variability in activity of CYP450 isoforms relevant to DEET metabolism was reported. Interest in biomarkers of exposure for pesticides has led to the development of high-throughput analytical methods that can be applied to large sample sizes for epidemiological studies. A liquid chromatography–tandem mass spectrometry method has been developed for the analysis of DEET and other pesticides in human urine samples (Olsson et al., 2004). In a recent cross-sectional study of human urine derived from a representative sample of the U.S. population (n  1977), geometric mean concentrations of DEET could not be calculated because of a high proportion of results below the limit of analytical detection (CDC, 2005).

98.4.3  Deet and Interactions with Other Environmental Chemicals (Ethanol and Sunscreens) Some investigators have reported a possible interaction between DEET absorption and ethanol ingestion. In an in vitro dermal penetration experiment, a single episode of oral ethanol consumption was reported to increase absorption of topically applied DEET (Brand et al., 2006). The clinical relevance of these findings is unclear, as in vivo studies of this potential interaction have not been reported. The mechanism through which this interaction may occur is uncertain, although vasodilation from ethanol consumption does not appear to be a likely explanation (Brand et al., 2007).

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N,N-diethyl-m-toluamide (DEET) CH3 CH2 N CH3

CH3 CH2

O

CYP2B6 CYP1A2

OXIDATION

CYP2C19 CYP3A4 CYP3A5 CYP2A6

DEALKYLATION

CH3 CH2 N CH2

CH3 CH2

H N

CH3

CH3 CH2

OH

O

N,N-diethyl-m-hydroxymethylbenzamide (BALC)

O

N-ethyl-m-toluamide (ET)

Figure 98.2  Metabolic pathways for DEET.

In recent years, several investigators have explored the combined use of sunscreen with insect repellents containing DEET. An in vitro diffusion study using piglet skin and poly (dimethylsiloxane) membrane reported that the permeation of DEET was synergistically enhanced when applied with the sunscreen oxybenzone (Gu et al., 2004). In this investigation, the observed effects were dependent upon the solvent vehicle and the concentrations of DEET and oxybenzone. Another in vitro experimental study in a hairless mouse skin model reported that the application of a commercial formulation of DEET with sunscreen (octocrylene, octylmethoxycinnamate, and benzophenone-3) resulted in more rapid detection of DEET using real-time mass spectroscopy (Ross et al., 2004). The role of chemicals other than sunscreens in enhancing the rate of dermal absorption from these formulations could not be excluded by the investigators. Interestingly, this study found no measurable detection of dermal penetration by any of the sunscreen compounds in the formulation mixture. A more recent in vivo study of piglet skin examined the effect of combined exposure to commercial formulations containing DEET and sunscreen (Kasichayanula et al., 2007). In comparison with a commercial formulation without sunscreen, the combined use of DEET and oxybenzone significantly enhanced percutaneous absorption and area under the curve measurements. While this study did not examine adverse effect endpoints, in combination

with the results of in vitro studies these findings suggest a need for further investigation to assess the efficacy as well as the safety of DEET when formulated or applied in combination with sunscreen.

98.4.4  Deet and Interactions with Other Environmental Chemicals The metabolism of DEET, when administered in combination with other environmental chemicals, has been the subject of several recent studies and reviews (Cho et al., 2007; Hodgson, 2008). These investigations have generally been based upon in vitro experiments using human hepatic microsomes. One study reported that the metabolism of DEET is inhibited by the organophosphate chlorpyrifos (Usmani et al., 2002). The same study also reported that preincubation of human microsomes with chlorpyrifos, permethrin, and pyridostigmine bromide alone or in combination can have stimulatory or inhibitory effects on DEET metabolism. These in vitro metabolic interactions have led to hypotheses about the potential role of combined exposure to DEET and other environmental chemicals in the etiology of Gulf War Syndrome. In one recent in vitro study of human liver microsomes and plasma incubations, the authors reported diminished DEET metabolism as a result of combined exposure with pyridostigmine bromide and

Chapter | 98  DEET

permethrin (Abu-Qare and Abou-Donia, 2008). The authors concluded that the results of this study support that combined exposure to these chemicals increases their neurotoxicity and explain the symptoms reported by Gulf War personnel. Well-designed clinical studies in human subjects exposed under realistic conditions do not provide support for this conclusion. In a randomized, double-blinded, placebo controlled trial to assess safety of exposure to DEET in combination with pyridostigmine, permethrin, and stress, 64 human volunteer subjects were evaluated for physiological endpoints, neurocognitive performance, and self-reported adverse effects (Roy et al., 2006). The results of this study found that short-term, repeated exposure to combined treatment was well tolerated and did not impair physical or neurocognitive performance. DEET was consistently detected in blood samples of exposed subjects, and no significant differences were observed in the Cmax, Tmax, or area under the curve by gender, weight, body mass index, or stress exposure.

98.4.5  Human Aspects: Clinical Case Reports (Dermal Reactions) Reports of significant dermal reactions have been infrequently reported from the normal use of DEET. Some case reports have described erythematous and bullous skin reactions occurring in healthy adults (military personnel) in association with the dermal application of DEET under conditions of occlusion. Consistent findings in these reports include the use of repellents with a high concentration of DEET (greater than 50%), and the onset of an erythematous rash within 24 h of application in certain treated areas of the skin (including the antecubital fossa and other regions where the skin may be occluded or covered by clothing during sleep) (Lamberg and Mulrennan, 1969; Reuveni and Yagupsky, 1982). In some of these case reports, the erythema evolved into a bullous eruption that subsequently ruptured, leaving an ulceration. In one case series, standard 48-h patch testing with a dilute formulation of DEET (5%) was not effective in confirming the diagnosis, although under skin occlusion conditions with a 75% DEET formulation a similar constellation of skin findings was observed in some individuals (Lamberg and Mulrennan, 1969). No other abnormal clinical signs or symptoms were reported in these cases. While the results of these case series are suggestive of an irritant contact dermatitis from repellents containing a high concentration of DEET under occlusion, such exposure conditions are unlike what are typically encountered during consumer use of insect repellents. Case reports of immediate hypersensitivity reactions in association with the dermal application of DEET have also been rarely reported (Maibach and Johnson, 1975; Miller, 1982).

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98.4.6  Clinical Case Reports: Adverse Neurological Effects Reports of adverse neurological reactions have rarely been reported in association with the dermal application of DEET (Osimitz and Murphy, 1997). A case of encephalopathy in a previously healthy 18-month-old male was described, which was temporally associated with the topical use of a repellent containing 17.6% DEET (Briassoulis et al., 2001). Infectious and metabolic etiologies were considered and ruled out based upon laboratory analyses. The child required intubation and had an uneventful recovery. In this case, details about the nature of the application (areas applied to skin, frequency of application) were not provided, making it difficult to interpret whether the effects were associated with misapplication or appropriate use of the formulation. The authors of this case report used logistic regression analysis to analyze other reported incidents of serious outcomes (fatalities) in children from dermal application of DEET. No association was observed between the concentration of DEET in the formulation, the duration of skin exposure, or the frequency of use (Briassoulis et al., 2001). A case of acute encephalopathy in a previously healthy adult was reported in association with the repeated application of insect repellants containing DEET (Hampers et al., 1999). In this case, the individual presented to an emergency department with acute onset of confusion and combativeness. The symptoms were temporally associated with the repeated application of two different insect repellants containing DEET (25% and 20%) on a hot summer day. One of the products was formulated as a combined sunscreen and insect repellant. Clinical signs included tremulousness, disorientation, combativeness, and generalized hypertonia. There was no evidence of hyperthermia, and an evaluation for infectious, metabolic, and other toxic etiologies (including ethanol) was unremarkable. The individual required aggressive supportive management, including intubation, and an uneventful neurological recovery followed. A serum specimen collected 16 h after presentation revealed a DEET concentration of 1.6 g/ml. Adverse neurological effects have also been reported in association with accidental and intentional ingestion exposures to DEET (Briassoulis et al., 2001; Tenenbein, 1987). In some of these incidents, co-ingestion of ethanol in the formulation was confirmed based upon blood analyses (Tenenbein, 1987), which may have influenced some of the neurological symptoms reported. Fatal outcomes in this case series were reported in association with incidents involving co-ingestants (including ethanol, hydralazine, and chlorpromazine).

98.4.6.1  Poison Control Center Data and DEET Exposure Incidents The first review of human exposure data for DEET gathered from Poison Control Centers (PCCs) in the United States, included over 9000 calls to PCCs reporting DEET exposures

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from 1985 to 1989 (Veltri et al., 1994). Approximately twothirds of the callers reported either no symptoms or only minor symptoms, such as skin or eye irritation that resolved rapidly. The authors concluded that the risk of serious medical effects with the labeled use of DEET-containing insect repellents was quite low. A subsequent publication based upon Poison Control Center data, including more than 20,000 DEET-related exposures during 1993–1997, reported similar conclusions (Bell et al., 2002). The majority of exposures were accidental and occurred at the subject’s home. Children experienced the most exposures, whereas teens experienced the least. Nearly three-fourths of the cases reported no symptoms related to the exposure. There was no apparent relationship between DEET concentration and the presence or severity of clinical signs and symptoms. Information routinely collected by PCCs is generally useful as a bellwether for safety issues for a drug or consumer product. However, there are limitations to these data (Bell et al., 2002; Veltri et al., 1994). Because details about the temporal association, prior medical history, and alternate etiologies are frequently lacking, it is difficult to accurately infer causality. For this reason, the National Registry of Human Exposure to DEET (DEET Registry) was created in 1995 by the DJV to evaluate the role of DEET, if any, in the more serious medical events that were reported. The Registry ran from 1995 to 2001. The objective was to identify individuals who reported serious symptoms (including seizures), obtain as much information as possible about the exposure and the symptoms, and follow those cases for up to 1 year. An important Registry feature was the estimation of the degree of causal relationship of the symptoms reported and exposure to DEET. The causality ratings included: possibly, probably, unlikely, unrelated, and relationship undetermined. Among the principles used in this assessment were the Bradford Hill Criteria: temporal relationship between exposure and reported effect, plausible etiological alternatives, consistency with previously reported symptoms, previous experience, and result of subsequent use of the product. Of 392 cases screened for inclusion in the Registry, 296 cases met the criteria for subsequent analysis. Seventy-six of these cases (19.4%) were not included because they typically had a minor severity rating or included dermal symptoms without any neurological or systemic component. Of these, 54 (18.2%) cases were judged to be unrelated or unlikely to be related to DEET exposure. The remaining 242 (81.8%) cases were determined to be either probably related (36 cases; 12.2%), possibly related (157 cases; 64.9%), or the causality was undetermined (49 cases; 16.6%) due to insufficient information. A total of 59 cases with seizures were reported, with 53 (90%) of these being of major or moderate severity (severity was not determined for the remaining six cases). During the 7-year history of the Registry, no seizure cases of “major” severity were judged to be “probably” related to

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DEET exposure. Only 12 cases (5.0%) had a major severity outcome rating whose symptoms were possibly related to exposure to DEET, and only six of these cases (2.5% of the Registry cases) were reported seizures. An additional five cases with major outcome were reported for which the causation could not be determined. Because there was no reliable estimate of the proportion of the DEET user population who use products with various DEET concentrations, it was impossible to determine if there is relationship between case severity and DEET concentration. No pattern was apparent in the Registry data to suggest an interaction of DEET with other products, such as overthe-counter or prescription medications. Despite the limitations in the use of passively reported data and considering the number of annual applications of DEET in the United States, relatively few reports of neurological adversity were documented in the Registry. The authors concluded that based on the information collected during the 7 years covered by the DEET Registry “it appears that the risk of a clinically significant adverse event from using DEET is very low.”

98.4.7  Regulatory Risk Assessment In 1998, the U.S. EPA completed a reregistration eligibility decision (RED) for DEET (U.S. EPA, 1998). Based upon oral and dermal LD50 studies, DEET was classified as low toxicity (Category III). DEET was classified as very low toxicity by inhalation (Category IV). Dermal sensitization was not apparent, and DEET was classified as low toxicity and very low toxicity for eye and dermal irritation, respectively. The U.S. EPA RED mutagenicity assays for DEET (including the Ames Assay, chromosomal aberration assay, and unscheduled DNA synthesis assay) were negative, and chronic toxicity studies on DEET did not show evidence of carcinogenicity. Based upon developmental and reproductive toxicity data in animal studies, the EPA concluded that the evidence did not support that DEET is uniquely toxic to infants and/or children. Neurotoxic effects were given special consideration in the risk assessment decision because of incident reports describing seizures and adverse neurological effects. The U.S. EPA reviewed incidents of adverse effects in humans and determined that it was not possible to conclude whether reports of seizures were directly related to DEET exposure. The FIFRA Scientific Advisory Panel (SAP) recommended that the Agency continue to collect and analyze incident reports relating to DEET (U.S. EPA, 1998). The U.S. EPA RED concluded that insect repellants containing DEET would generally not cause unreasonable risks to humans or the environment. Additional label warnings and restrictions were required for repellents containing DEET. Child safety claims were prohibited on the label, and all insect repellants containing DEET were required to include additional statements intended to be protective for children and other sensitive individuals.

Chapter | 98  DEET

The Pest Management Regulatory Agency in Canada has also completed a reevaluation decision document on insect repellants containing DEET (Pest Management Regulatory Agency, 2002). The risk assessment process included calculation of estimated margins of exposure (MOE) for both acute and intermediate-term dermal exposure to DEET in infants, children, and adults. The intermediate-term MOEs for products with greater than 30% DEET were less than 100 for adults. For children 2–12 years of age, products with greater than 10% DEET had MOEs less than 100. For toddlers, acute (single application) MOEs were only acceptable for products with up to 10% DEET, whereas intermediateterm MOEs were inadequate for all DEET concentrations. Based upon the risk assessment, the Pest Management Regulatory Agency in Canada decided that products containing more than 30% DEET were no longer acceptable for registration, and that products containing no more than 30% DEET would provide adults and individuals 12 years and older with sufficient protection. For children between 2 and 12 years of age, the regulatory agency recommended that the least concentrated solution (10% or less) of DEET be used, with no more than three applications per day. For children aged 6 months to 2 years, where a high risk of complications from insect bites exists, the use of one application per day may be considered, with the least concentrated product (less than 10%) of DEET to be used. The agency decision was to not use insect repellants containing DEET on children less than 6 months of age.

98.4.7.1  Conclusions of Other Authorities In addition to regulatory authorities, other professional groups have issued recommendations on using DEET safely. The American Academy of Pediatrics has advised that insect repellents containing DEET with a concentration of 10% are as safe as products containing up to 30% when used according to the labeling instructions (AAP Committee on Environmental Health, 2003). The Academy advised against the use of DEET on children under 2 months of age. The Academy also advised that DEET should not be used in products that combine the repellent with a sunscreen (AAP Committee on Environmental Health, 2003). Other authorities have concluded that DEET is highly effective and generally considered safe when used according to the product labeling (Koren et al., 2003; Roberts and Reigart, 2004). A recently published risk assessment for DEET found no significant toxicological risks from typical patterns of use (Antwi et al., 2008).

References AAP Committee on Environmental Health (2003). Follow safety precautions when using DEET on children. AAP News 22, 99. Abdel-Rahman, A., Shetty, A. K., and Abou-Donia, M. B. (2001). Subchronic dermal application of N,N-diethyl m-toluamide (DEET)

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and permethrin to adult rats, alone or in combination, causes diffuse neuronal cell death and cytoskeletal abnormalities in the cerebral cortex and the hippocampus, and Purkinje neuron loss in the cerebellum. Exp. Neurol. 172, 153–171. Abou-Donia, M. B., Wilmarth, K. R., Jensen, K. F., Oehme, F. W., and Kurt, T. L. (1996). Neurotoxicity resulting from coexposure to pyridostigmine bromide, deet, and permethrin: implications of Gulf War chemical exposures. J. Toxicol. Environ. Health 48, 35–56. Abou-Donia, M. B., Goldstein, L. B., Dechovskaia, A., Bullman, S., Jones, K. H., Herrick, E. A., Abdel-Rahman, A. A., and Khan, W. A. (2001). Effects of daily dermal application of DEET and epermethrin, alone and in combination, on sensorimotor performance, blood-brain barrier, and blood-testis barrier in rats. J. Toxicol. Environ. Health A 62, 523–541. Abu-Qare, A. W., and Abou-Donia, M. B. (2008). In vitro metabolism and interactions of pyridostigmine bromide, N,N-diethyl-m-toluamide, and permethrin in human plasma and liver microsomal enzymes. Xenobiotica 38, 294–313. Antwi, F. B., Shama, L. M., and Peterson, R. K. (2008). Risk assessments for the insect repellents DEET and picaridin. Regulat. Toxicol. Pharmacol. 51, 31–36. Bell, J. W., Veltri, J. C., and Page, B. C. (2002). Human exposures to N,N-diethyl-m-toluamide insect repellents reported to the American Association of Poison Control Centers 1993-1997. Int. J. Toxicol. 21, 341–352. Blomquist, L., and Thorsell, W. (1977). Distribution and fate of the insect repellent 14C-N, N-diethyl-m-toluamide in the animal body. II. Distribution and excretion after cutaneous application. Acta Pharmacol. Toxicol. 41, 235–243. Brand, R. M., Jendrzejewski, J. L., Henery, E. M., and Charron, A. R. (2006). A single oral dose of ethanol can alter transdermal absorption of topically applied chemicals in rats. Toxicol. Sci. 92, 349–355. Brand, R. M., Jendrzejewski, J. L., and Charron, A. R. (2007). Potential mechanisms by which a single drink of alcohol can increase transdermal absorption of topically applied chemicals. Toxicology 235, 141–149. Briassoulis, G., Narlioglou, M., and Hatzis, T. (2001). Toxic encephalo­ pathy associated with use of DEET insect repellents: a case analysis of its toxicity in children. Hum. Exp. Toxicol. 20, 8–14. Brown, M., and Hebert, A. A. (1997). Insect repellents: an overview. J. Am. Acad. Dermatol. 36, 243–249. Brusick, D. (1976). “Mutagenicity Evaluation of m-DEET”. Unpublished study conducted at Litton Bionetics, Inc. under the sponsorship of McLaughlin Gormley King Co. Brusick, D. (1980). “The Effect of DEET (N,N-diethyltoluamide) on the Morphology, Viability, and Count of Sperm in Rats Exposed by Dermal Administration”. Unpublished study conducted at Litton Bionetics, Inc. under the sponsorship of McLaughlin Gormley King Co. Centers for Disease Control and Prevention (CDC). (2005). “Third National Report on Human Exposure to Environmental Chemicals”. Atlanta, Georgia. Centers for Disease Control and Prevention (CDC) (2008). Chapter 2: Pre- and post-travel general health recommendations. In “CDC Health Information for International Travel” (P. E. Kozarksky, P. M. Arguin, and C. Reed, eds.). Elsevier, Philadelphia. Chaney, L. A., Rockhold, R. W., Wineman, R. W., and Hume, A. S. (1999). Anticonvulsant-resistant seizures following pyridostigmine bromide (PB) and N,N-diethyl-m-toluamide (DEET). Toxicol. Sci. 49, 306–311. Cho, T. M., Rose, R. L., and Hodgson, E. (2007). The effect of chlorpyrifos-oxon and other xenobiotics on the human cytochrome P450-dependent metabolism of naphthalene and deet. Drug Metabol. Drug Interact. 22, 235–262.

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Committee to Advise on Tropical Medicine and Travel (CATMAT) (2005). Statement on personal protective measures to prevent arthropod bites – Update. Can. Commun. Dis. Rep. 31, 1–18. Defense Pest Management Information Analysis Center (DPMIAC) (2002). “Technical Guide No. 36: Personal Protective Measures Against Insects and Other Arthropods of Military Significance”, pp. 21–25. Defense Pest Management Information Analysis Center (DPMIAC), Armed Forces Pest Management Board (AFPMB), Washington, DC. Dogan, E. B., Ayres, J. W., and Rossignol, P. A. (1999). Behavioural mode of action of deet: inhibition of lactic acid attraction. Med. Vet. Entomol. 13, 97–100. Domb, A. J., Marlinsky, A., Maniar, M., and Teomim, L. (1995). Insect repellent formulations of N,N-diethyl-m-toluamide (deet) in a liposphere system: efficacy and skin uptake. J. Am. Mosq. Control Assoc. 11, 29–34. Dorman, D. C., Buck, W. B., Trammel, H. L., Jones, R. D., and Beasley, V. R. (1990). Fenvalerate/N,N-diethyl-m-toluamide (Deet) toxicosis in two cats. J. Am. Vet. Med. Assoc. 196, 100–102. Feldmann, R. J., and Maibach, H. I. (1970). Absorption of some organic compounds through the skin in man. J. Invest. Dermatol. 54, 399–404. Fradin, M. S., and Day, J. F. (2002). Comparative efficacy of insect repellents against mosquito bites. N. Engl. J. Med. 347, 13–18. Goldenthal, E. (1989a). “Evaluation of DEET in a 90-Day Dermal Toxicity Study in Castrated Male Rats”. Unpublished study conducted at International Research and Development Corporation under the sponsorship of the DEET Joint Venture/Chemical Specialties Manufacturers Association. Goldenthal, E. (1989b). “Evaluation of DEET in a 90-Day Dose Range Findings Study in Hamsters”. Unpublished study conducted at International Research and Development Corporation under the sponsorship of the DEET Joint Venture/Chemical Specialties Manufacturers Association. Goldenthal, E. (1991). “Evaluation of DEET in a 90-Day Subchronic Dermal Toxicity Study in Micropigs®”. Unpublished study conducted at International Research and Development Corporation under the sponsorship of the DEET Joint Venture/Chemical Specialties Manufacturers Association. Goldenthal, E. (1992). “Evaluation of DEET in a Multistrain 90-Day Dietary Renal Toxicity Study in Rats”. Unpublished study conducted at International Research and Development Corporation under the sponsorship of the DEET Joint Venture/Chemical Specialties Manufacturers Association. Gu, X., Kasichayanula, S., Fediuk, D. J., and Burczynski, F. J. (2004). In-vitro permeation of the insect repellent N,N-diethyl-m-toluamide (DEET) and the sunscreen oxybenzone. J Pharm. Pharmacol. 56, 621–628. Gujral, I. B., Zielinski-Gutierrez, E. C., LeBailly, A., and Nasci, R. (2007). Behavioral risks for West Nile virus disease, northern Colorado, 2003. Emerg. Infect. Dis. 13, 419–425. Hampers, L. C., Oker, E., and Leikin, J. B. (1999). Topical use of DEET insect repellent as a cause of severe encephalopathy in a healthy adult male. Acad. Emerg. Med. 6, 1295–1297. Hoffmann, E. J., and Miller, J. R. (2003). Reassessment of the role and utility of wind in suppression of mosquito (Diptera: Culicidae) host finding: stimulus dilution supported over flight limitation. J. Med. Entomol. 40, 607–614. Hodgson, E., and Rose, R. L. (2008). Metabolic interactions of agrochemicals in humans. Pest Manage. Sci. 64, 617–621. Iscan, Y., Hekimoglu, S., Sargon, M. F., and Hincal, A. A. (2006). DEETloaded solid lipid particles for skin delivery: In vitro release and skin

Hayes’ Handbook of Pesticide Toxicology

permeation characteristics in different vehicles. J. Microencapsul. 23, 315–327. Johnson, D. (1987a). “Evaluation of DEET in a 90-Day Oral Dose Range Finding Study in Mice”. Unpublished study conducted at International Research and Development Corporation under the sponsorship of the DEET Joint Venture/Chemical Specialties Manufacturers Association. Johnson, D. (1987b). “Evaluation of DEET in a 90-Day Oral Dose Range Finding Study in Rats”. Unpublished study conducted at International Research and Development Corporation under the sponsorship of the DEET Joint Venture/Chemical Specialties Manufacturers Association. Johnson, D. (1987c). “Evaluation of DEET in a 90-Day Subchronic Dermal Toxicity Study in Rats”. Unpublished study conducted at International Research and Development Corporation under the sponsorship of the DEET Joint Venture/Chemical Specialties Manufacturers Association. Jortner, B. S. (2006). The return of the dark neuron. A histological artifact complicating contemporary neurotoxicologic evaluation. NeuroToxicology 27, 628–634. Kasichayanula, S., House, J. D., Wang, T., and Gu, X. (2007). Percutaneous characterization of the insect repellent DEET and the sunscreen oxybenzone from topical skin application. Toxicol. Appl. Pharmacol. 223, 187–194. Kasting, G. B., Bhatt, V. D., and Speaker, T. J. (2008). Microencapsulation decreases the skin absorption of N,N-diethyl-m-toluamide (DEET). Toxicol. in Vitro 22, 548–552. Koren, G., Matsui, D., and Bailey, B. (2003). DEET-based insect repellents: safety implications for children and pregnant and lactating women. Can. Med. Assoc. J. 169, 209–212. Lamberg, S. I., and Mulrennan, J. A. Jr. (1969). Bullous reaction to diethyl toluamide (DEET). Resembling a blistering insect eruption. Arch. Dermatol. 100, 582–586. Macko, J., and Bergman, J. (1980). “Phase 4-Inhalation Toxicities of N,NDiethyl-meta-toluamide (M-Det)”. Unpublished study conducted at U.S. Army Environmental Hygiene Agency. Macko, J., and Weeks, M. (1980). “Acute Toxicity Evaluation of N,NDiethyl-meta-toluamide (M-Det)”. Unpublished study conducted at U.S. Army Environmental Hygiene Agency under the sponsorship of S.C. Johnson & Son, Inc. Maibach, H. I., and Johnson, H. L. (1975). Contact urticaria syndrome. Contact urticaria to diethyltoluamide (immediate-type hypersensitivity). Arch. Dermatol. 111, 726–730. McCain, W. C., Lee, R., Johnson, M. S., Whaley, J. E., Ferguson, J. W., Beall, P., and Leach, G. (1997). Acute oral toxicity study of pyridostigmine bromide, permethrin, and DEET in the laboratory rat. J. Toxicol. Environ. Health 50, 113–124. Miller, J. D. (1982). Anaphylaxis associated with insect repellent. N. Engl. J. Med. 307, 1341–1342. Moody, R. P., and Maibach, H. I. (2006). Skin decontamination: Importance of the wash-in effect. Food Chem. Toxicol. 44, 1783–1788. Moody, R. P., Nadeau, B., and Chu, I. (1995). In vitro dermal absorption of N,N-diethyl-m-toluamide (DEET) in rat, guinea pig, and human skin. In Vitro Toxicol. 8, 263–275. Moore, G. (2000). “Acute Oral, Dermal and Inhalation Toxicity, Primary Skin and Eye Irritation and Dermal Sensitization Studies with DEET Insect Repellent”. Unpublished studies conducted at Product Safety Labs under the sponsorship of Morflex, Inc. Mount, M. E., Moller, G., Cook, J., Holstege, D. M., Richardson, E. R., and Ardans, A. (1991). Clinical illness associated with a commercial tick and flea product in dogs and cats. Vet. Hum. Toxicol. 33, 19–27. Olsson, A. O., Baker, S. E., Nguyen, J. V., Romanoff, L. C., Udunka, S. O., Walker, R. D., Flemmen, K. L., and Barr, D. B. (2004). A liquid chroma-

Chapter | 98  DEET

tography–tandem mass spectrometry multiresidue method for quantification of specific metabolites of organophosphorus pesticides, synthetic pyrethroids, selected herbicides, and deet in human urine. Anal. Chem. 76, 2453–2461. Osimitz, T. G., and Murphy, J. V. (1997). Neurological effects associated with use of the insect repellent N,N-diethyl-m-toluamide (DEET). J. Toxicol. Clin. Toxicol. 35, 435–441. Pest Management Regulatory Agency (2002). “Re-evaluation Decision Document: Personal Insect Repellents Containing DEET (N,Ndiethyl-m-toluamide and Related Compounds)”. RRD2002-01. Qiu, H., Jun, H. W., Dzimianski, M., and McCall, J. (1997). Reduced transdermal absorption of N,N-diethyl-m-toluamide from a new topical insect repellent formulation. Pharm. Dev. Technol. 2, 33–42. Reuveni, H., and Yagupsky, P. (1982). Diethyltoluamide-containing insect repellent: adverse effects in worldwide use. Arch. Dermatol. 118, 582–583. Roberts, J. R., and Reigart, J. R. (2004). Does anything beat DEET? Pediatr. Ann. 33, 443–453. Ross, E. A., Savage, K. A., Utley, L. J., and Tebbett, I. R. (2004). Insect repellant interactions: sunscreens enhance deet (n,n-diethyl-m-toluamide) absorption. Drug Metab. Dispos. 32, 783–785. Rowland, M., Freeman, T., Downey, G., Hadi, A., and Saeed, M. (2004). DEET mosquito repellent sold through social marketing provides personal protection against malaria in an area of all-night mosquito biting and partial coverage of insecticide-treated nets: a case-control study of effectiveness. Trop. Med. Int. Health 9, 343–350. Roy, M. J., Kraus, P. L., Seegers, C. A., Young, S. Y., Kamens, D. R., Law, W. A., Cherstniakova, S. A., Chang, D. N., Cooper, J. A., Sato, P. A., Matulich, W., Krantz, D. S., Cantilena, L. R., and Deuster, P. A. (2006). Pyridostigmine, diethyltoluamide, permethrin, and stress: A doubleblind, randomized, placebo-controlled trial to assess safety. Mayo Clin. Proc. 81, 1303–1310. San, R., and Schadly, M. (1989). “Salmonella/Mammalian-Microsome Plate Incorporation Mutagencity Assay (Ames Test) with a Confirmatory Assay”. Unpublished study conducted at Microbiological Associates, Inc. under the sponsorship of the DEET Joint Venture/Chemical Specialties Manufacturers Association. Schardein, J. (1989). “Evaluation of DEET in a Two Generation Reproduction/Fertility Study in Rats”. Unpublished study conducted at International Research and Development Corporation under the sponsorship of the DEET Joint Venture/Chemical Manufacturers Association. Schoenig, G. P. (2002). Locomotor and sensorimotor performance deficit in rats following exposure to pyridostigmine bromide, DEET, and permethrin. Alone and in combination. Toxicol. Sci. 68, 516–517. Schoenig, G. P., Hartnagel, R. E. Jr., Schardein, J. L., and Vorhees, C. V. (1993). Neurotoxicity evaluation of N,N-diethyl-m-toluamide (DEET) in rats. Fundam. Appl. Toxicol. 21, 355–365. Schoenig, G. P., Neeper-Bradley, T. L., Fisher, L. C., and Hartnagel, R. E. Jr. (1994). Teratologic evaluations of N,N-diethyl-m-toluamide (DEET) in rats and rabbits. Fundam. Appl. Toxicol. 23, 63–69. Schoenig, G. P., Hartnagel, R. E. Jr., Osimitz, T. G., and Llanso, S. (1996). Absorption, distribution, metabolism, and excretion of N,Ndiethyl-M-toluamide in the rat. Drug Metab. Dispos. 24, 156–163.

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Schoenig, G. P., Osimitz, T. G., Gabriel, K. L., Hartnagel, R., Gill, M. W., and Goldenthal, E. I. (1999). Evaluation of the chronic toxicity and oncogenicity of N,N-diethyl-m-toluamide (DEET). Toxicol. Sci. 47, 99–109. Selim, S., Hartnagel, R. E. Jr., Osimitz, T. G., Gabriel, K. L., and Schoenig, G. P. (1995). Absorption, metabolism, and excretion of N,N-diethyl-m-toluamide following dermal application to human volunteers. Fundam. Appl. Toxicol. 25, 95–100. Smallwood, A. W., DeBord, K. E., and Lowry, L. K. (1992). N,N-diethylm-toluamide (m-DET): analysis of an insect repellent in human urine and serum by high-performance liquid chromatography. J. Anal. Toxicol. 16, 10–13. Stenbäck, F. (1976). “Testing of Cosmetics, Ingredients of Sunscreen ointments, Insect Repellents, and Detergents on Skin of Mice and Rabbits: Lifespan Studies”. Unpublished study conducted at the Eppley Institute for Research in Cancer and Allied Diseases under the sponsorship of Morflex Chemical Co., Inc. Stinecipher, J., and Shah, J. (1997). Percutaneous permeation of N,Ndiethyl-m-toluamide (DEET) from commercial mosquito repellents and the effect of solvent. J. Toxicol. Environ. Health 52, 119–135. Swentzel, K. (1978). “Investigation of N-N-Diethyl-m-Toluamide (M-Det) for Dominant Lethal Effects in the Mouse Study”. Unpublished study conducted at U.S. Army Environmental Hygiene Agency. Syed, Z., and Leal, W. S. (2008). From the cover: Mosquitoes smell and avoid the insect repellent DEET. Proc. Natl. Acad. Sci. 105, 13598–13603. Tenenbein, M. (1987). Severe toxic reactions and death following the ingestion of diethyltoluamide-containing insect repellents. JAMA 258, 1509–1511. U.S. EPA (1998). “Reregistration Eligibility Decision: DEET”. Office of Pesticide Programs. U.S. Environmental Protection Agency, Washington, DC. Usmani, K. A., Rose, R. L., Goldstein, J. A., Taylor, W. G., Brimfield, A. A., and Hodgson, E. (2002). In vitro human metabolism and interactions of repellent N,N-diethyl-m-toluamide. Drug Metab. Dispos. 30, 289–294. Veltri, J. C., Osimitz, T. G., Bradford, D. C., and Page, B. C. (1994). Retrospective analysis of calls to poison control centers resulting from exposure to the insect repellent N,N-diethyl-m-toluamide (DEET) from 1985–1989. J. Toxicol. Clin.Toxicol. 32, 1–16. Verschoyle, R. D., Brown, A. W., Nolan, C., Ray, D. E., and Lister, T. (1992). A comparison of the acute toxicity, neuropathology, and electrophysiology of N,N-diethyl-m-toluamide and N,N-dimethyl-2,2diphenylacetamide in rats. Fundam. Appl. Toxicol. 18, 79–88. Weil, C. (1973). “N-N-Diethyl-m-Toluamide Range Finding Toxicity Studies”. Unpublished studies conducted at Carnegie-Mellon University under the sponsorship of Union Carbide Corporation. Wright, D. M., Hardin, B. D., Goad, P. W., and Chrislip, D. W. (1992). Reproductive and developmental toxicity of N,N-diethyl-m-toluamide in rats. Fundam. Appl. Toxicol. 19, 33–42. Zeiger, E., Anderson, B., Haworth, S., Lawlor, T., and Mortelmans, K. (1992). Salmonella mutagenicity tests: V. Results from the testing of 311 chemicals. Environ. Mol. Mutagen 19(Suppl 21), 2–141.

Chapter 99

The Safety Assessment of Piperonyl Butoxide Thomas G. Osimitz Science Strategies, LLC

99.1  Chemistry and formulations Piperonyl butoxide (PBO), 2-(2-butoxyethoxy)ethyl 6-propyl-piperonyl ether (IUPAC), is an insecticide synergist produced from the condensation of the sodium salt of 2-(2-butoxyethoxy) ethanol and the chloromethyl derivative of hydrogenated safrole (dihydrosafrole) (Figure 99.1). The dihydrosafrole moiety containing the methylenedioxyphenyl ring constitutes over half of the PBO molecule by weight and is traditionally derived from sassafras oil. Sassafras oil is an essential oil distilled from ­ several species of trees found in China, Laos, and Burma. However, the availability of sassafras oil is limited today due to environmental concerns related to the destructive harvesting of trees, as well as to the nature of safrole as a drug precursor. A novel manufacturing process for PBO where dihydrosafrole is derived no longer from Sassafras oil but from catechol was patented first in 1998 by Endura S.p.A (Endura, 2009). Current world production of PBO averages 94% pure. In the early days, PBO contained small but detectable amounts of safrole and dihydrosafrole (DHS). However, refinements in distillation have resulted in safrole and DHS levels usually below the 40 ppm detection limit by highresolution gas chromatography (Di Blasi, 1998). When the PBO precursor DHS is manufactured from catechol, PBO does not contain safrole as an impurity.

O O

H2 C

C H2

C H2 O

CH3

C H2

H2

H2

C

O

O

Figure 99.1  Chemical structure of PBO. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

H2 C H2

O

C H2

C

C H2

CH3

The development of PBO grew out of a need in the late 1930s and early 1940s to extend the usefulness of the naturally derived insecticide pyrethrum, which was considered a strategic insecticide against mosquitoes and other disease-carrying insects. The chemicals that were developed had little intrinsic pesticidal activity of their own; however, they did increase the effectiveness of a given dose of pyrethrins and were thus called synergists. PBO was one of a series of molecules synthesized (Wachs, 1947). PBO is usually formulated with natural pyrethrins or synthetic pyrethroids in ratios (PBO: pyrethrins) ranging from 3:1 to 20:1. Formulations of PBO and carbamates are also available, although their use is minor relative to that of PBO and pyrethrins/pyrethroids.

99.2  Uses As a synergist, PBO inhibits the mixed function oxidase (MFO) system of insects, thereby reducing the oxidative breakdown of other pesticides such as pyrethrum and the synthetic pyrethroids (Casida, 1970). The precise mechanism of inhibition is unknown, but speculation is that a carbene derivative forms and binds to the heme moiety of the cytochrome P450 enzyme, thereby rendering it inactive (Dahl and Brenzinski, 1985; Delaforge et al., 1985; Franklin, 1976; Hodgson et al., 1973; Murray and Reidy, 1989; Philpot and Hodgson, 1971, 1972a,b). More recently, PBO has also been found to inhibit resistance-associated esterases in many insects (Khot et al., 2008) and to inhibit other enzymes such as hydrolases (Kakko et al., 2000) and glutathione S-transferases (Varsano et al., 1992). The result is that higher levels of the insecticide remain in the insect and are thereby available to exercise their lethal effect on the insect. PBO enhances the pesticidal activity of 2127

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Table 99.1  Summary of Acute Toxicity Data and Classifications Route

Species

Result

EEC labeling classification

USEPA toxicity category

Reference

Oral LD50

Rat

4 g/kg (male) 7 g/kg (female)

Unclassified

Category IV

Gabriel (1991b)

Dermal LD50

Rabbit

2 g/kg

Unclassified

Category IV

Gabriel (1991a)

Inhalation LC50

Rat

5.9 mg/l air

Harmful

Category III

Hoffman (1991)

Eye irritation

Rabbit

Minimally irritating

Labeling not indicated

Category III

Romanelli (1991b)

Skin irritation

Rabbit

Minimally irritating

Labeling not indicated

Category IV

Romanelli (1991c)

Skin sensitization (Buehler)

Guinea pig

Negative

Labeling not indicated

Category IV

Romanelli (1991a)

a given level of active ingredient, thus promoting reduced use of the pesticide. Appearing in over 1500 U.S. Environmental Protection Agency (U.S. EPA)-registered products, PBO is one of the most commonly registered pesticides in terms of the number of formulas in which it is present. It is approved for preharvest application to a wide variety of crops including fruits and vegetables. The application rates are low; the highest single rate is 0.5 lbs PBO/acre. It is also used extensively in combination with pyrethrins, various synthetic pyrethroids, and other insecticides to control insect pests in and around the home and in food-handling establishments. A wide variety of water-based PBO-containing products such as crack and crevice sprays, total release foggers, and flying insect sprays are made for use by consumers in the home. Annual use of PBO in the United States is approximately 1.3–1.5 million lbs (PBO Task Force, 2009). About 47% of this total is used for indoor residential purposes, 17% for indoor food uses in warehouses and food handling establishments, and only 13% goes for agricultural crop applications. PBO has also been allowed as a food additive in Japan since 1955, its maximum approved level being 0.024 g/kg (24 ppm) in raw cereals.

99.3  Hazard identification 99.3.1  Acute Toxicity Numerous acute studies have been conducted over the years with PBO in a variety of species and by various exposure routes. This body of data, including the most recent studies, indicates that PBO is generally of low acute toxicity to animals. It is mildly irritating to the eye and skin. It is not a dermal sensitizer. Table 99.1 summarizes this acute toxicity data as well as European Economic Commission (EEC) labeling classifications and U.S. EPA toxicity categories.

99.3.2  Subchronic Toxicity PBO has been tested in dogs, mice, rats, rabbits, and African green monkeys for subchronic toxicity. A summary of the subchronic toxicity studies discussed below is presented in Table 99.2.

99.3.2.1  Dogs Lorber (1972) reported unexpected alterations in blood counts of intact and splenectomized dogs after use of a fogger (containing PBO among other chemicals). The fogger bathed the animals inadvertently in a “dense pesticide mist.” The dogs were part of a research project investigating the relationship of spleen, bone marrow, and blood cells. Further work was undertaken in which 17 intact, 12 splenectomized, and seven partially splenectomized dogs were purposefully exposed to deodorized kerosene containing only PBO (1.5%). Exposure periods consisted of four intervals of 5 min duration, with an 8-min interval between each exposure. The splenectomized dogs showed a reduction in serum platelet count and an occasional increase in reticulocytes. The authors concluded that the demonstrated greater resistance of intact dogs to the hematotoxic potential of the tested chemicals may have been in part due to the larger spleens in these animals, which could perhaps sequester the chemical(s) more effectively. Moreover, they felt that the normal hepatic blood flow in these animals could also enhance the removal of the chemical(s) from the systemic circulation. Goldenthal (1993a) conducted a range-finding study as a prelude to a 1-year chronic study. PBO was administered in the diet to dogs (four animals/sex/dose level) for 8 weeks. The dosage levels were 500, 1000, 2000, and 3000 ppm (approximately equivalent to 12.5, 25, 50, and 75 mg/kg body weight/day, respectively). All dogs survived to study termination. All animals in the 3000 ppm dose group had decreased appetites and reduced defecation during the 1st week. No other abnormal clinical signs were

Chapter | 99  The Safety Assessment of Piperonyl Butoxide

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Table 99.2  Summary of Results of Subchronic Toxicity Studies with PBO Species Route Dog

Mouse

Rat

Rabbit

Dose

Inhalation 15,000 ppm

Duration NOAEL

Comments

Reference

Four 5-min intervals

Not applicable

Reduction in serum Lorber (1972) platelet count, increase in reticulocytes (splenectomized animals)

Diet

8 weeks 500–3000 ppm (12.5– 75 mg/kg body weight/day)

500 ppm (12.5 mg/ kg body weight/day)

Decreased body weight, increased liver weight, hepatocyte hypertrophy

Goldenthal (1993a)

Diet

1000–9000 ppm (150– 1350 mg/kg body weight/ day)

20 days

1000 ppm (150 mg/ kg body weight/day)

Increased liver weight, hepatocyte hypertrophy, necrosis, inflammatory cell infiltration

Fujitani et al. (1993)

Oral

10–1000 mg/kg body weight/day

90 days

30 mg/kg body weight/day

Increased liver weight, liver necrosis, centrilobular hypertrophy

Chun and Wagner (1993)

Oral

1500–6000 ppm (236– 880 mg/kg body weight/ day)

7 wks

Not established

Alterations in motor activity

Tanaka (1993)

Gavage

2.5–5 ml/kg body weight/ day

31 days

Not established

Anorexia, loss of weight, death

Sarles and Vandergrift (1952)

Oral

1857 mg/kg body weight/ day

90 days

Not applicable

40% mortality, increased liver Bond et al. (1973) weight

Diet

62.5–2000 mg/kg body weight/day

28 days

125 mg/kg body weight/day

Increased liver weight, microscopic changes in liver

Modeweg-Hausen et al. (1984)

Oral

6000–24,000 ppm (300–1200 mg/kg body weight/day)

13 weeks Not established

Decreased body weight, increased liver and kidney weights, hepatocyte hypertrophy

Fujitani et al. (1992)

Gavage

250–4000 mg/kg body weight/day

10 days

250 mg/kg body weight/day

Ataxia, twitching, dyspnea, gastric ulceration

Chun and Neeper-Bradley (1992)

Inhalation 15–512 mg/m3

90 days

155 mg/m3

Alterations in clinical chemistry Newton (1992) parameters, increased liver and kidney weights

Oral

1 or 4 ml/kg body weight/ wk (5% emulsion)

3 weeks

Not applicable

No signs of toxicity

Dermal

100–1000 mg/kg body weight/day

3 weeks

1000 mg/kg body weight/day (systemic toxicity) 100 mg/kg body weight/day (local effects)

Slight erythema/edema, Goldenthal (1992) fissuring/inflammation of skin

0.03 or 0.1 ml/kg body weight/day

4 weeks

Not applicable

Minor changes in liver

Monkey Oral

present. Three out of four dogs lost weight in the highest dose group. Even at 1000 ppm PBO in the diet, weight gains were lower than the control group. Food intake was similar between treated and control groups, except for a slight reduction in some animals at the 3000 ppm dosage level. There were no treatment-related effects on hematological parameters at any dose level, but slight increases in

Sarles et al. (1949)

Sarles and Vandergrift (1952)

alkaline phosphatase values and slight decreases in cholesterol were noted at doses 2000 ppm. No treatment-related microscopic changes were noted at necropsy in any group, but a compound-related increase in absolute and relative liver and gall bladder weights was recorded in males. Upon histopathologic examination, hypertrophy of hepatocytes was noted in males of all dose levels and in females at

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dosages 2000 ppm and above. This finding was consistent with the increases in the liver weights and serum alkaline phosphatase levels described above. No other ­ treatmentrelated microscopic changes were evident. There was a decrease in the absolute and relative weights of the testes and epididymis in the groups treated with 2000 and 3000 ppm. The dose level of 500 ppm was set as a noobservable-adverse-effect level (NOAEL) for this study because the changes recorded in the liver were considered adaptive in nature rather than adverse and were not accompanied by any systemic signs of toxicity.

99.3.2.2  Mice Fujitani et al. (1993) reported the results of dosing CD-1 mice (10 animals/sex/dose level) with 1000, 3000, or 9000 ppm (approximately equivalent to 150, 450, and 1350 mg/kg body weight/day, respectively) PBO in the diet for 20 days. Body weights were depressed in the high-dose animals (about 15%) and in the mid-dose females (about 8%). Kidney and spleen weights were also reduced in the high-dose group. A treatment-related elevation in liver weights was noted with a 79% increase in the high-dose males. Females in the high-dose group showed higher levels of -glutamyl transpeptidase (GGT) activity. The high-dose males and females featured higher levels of cholesterol, phospholipids, and total serum proteins. Hepatocyte hypertrophy, single cell necrosis, and inflammation, most prominently in the centrilobular region, were also seen in the livers of the mid- and high-dose groups. The NOAEL for this study, based on liver toxicity, was 1000 ppm. As a prelude to a 2-year bioassay, Chun and Wagner (1993) reported the conduct of a 90-day oral toxicity study in CD-1 mice (15 animals/sex/dose level) using dose levels of 10, 30, 100, 300, and 1000 mg PBO/kg body weight/ day in the diet. A significant decrease in body weight was noted in the high-dose males (34% versus controls). The target organ for toxicity was the liver as indicated by increased liver weights, hepatocyte necrosis, and centrilobular hypertrophy (NOAEL  30 mg/kg body weight/day).

99.3.2.3  Rats Sarles and Vandergrift (1952) gavaged six male and six female rats with PBO daily, 6 days/week for 31 days. The first seven doses were at 2.5 ml/kg body weight. Two animals died on the 3rd and 4th days. The remaining animals improved after some initial clinical signs of toxicity. They were dosed with a second round of seven doses of 3.5 ml/kg body weight each. Little toxicity was noted at this dose level. Hence, the animals thereafter received doses of 5 ml/kg body weight. Clinical signs of toxicity included anorexia and loss of weight. Additional animal deaths occurred from the 17th to 24th days of testing; the next and last death was at 31 days.

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Bond et al. (1973) administered 1857 mg PBO/kg body weight/day orally to a single group of 20 rats for 90 days. Forty percent of the rats died prior to conclusion of the study. The most significant finding was a dramatic increase in liver weight (i.e., 2.4 times that of untreated controls). The authors also alluded to another study they conducted, which is unpublished, in which rats were fed 500 mg PBO/kg body weight/day. These animals were reported to have liver and kidney damage. A 4-week range-finding study was conducted by Modeweg-Hausen et al. (1984). Rats (10 animals/sex/dose level) were fed 62.5, 125, 250, 500, 1000, or 2000 mg PBO/kg body weight/day. Hepatic eosinophilic infiltration and the increased vacuolization of hepatocytes were seen with increasing severity among the mid- and high-dose groups. These effects were viewed as being degenerative changes representing chronic toxicity. Liver weights were elevated at 250 mg/kg body weight/day and above in the males and at 500 mg/kg body weight/day and above in the females. Except for an increase in alkaline phosphatase at the highest dose level, no treatment-related changes were reported in hematologic and clinical chemical parameters. Based on the observed liver toxicity, the NOAEL for this study was 125 mg/kg body weight/day. A 13-week subacute oral toxicity study was conducted in Fischer F344 rats (10 animals/sex/dose level) at levels of 6000, 12,000, or 24,000 ppm (approximately equivalent to 300, 600, or 1200 mg/kg body weight/day, respectively) PBO in the diet (Fujitani et al., 1992). No mortality occurred. Nasal bleeding and dose-related abdominal distension were reported. A significant decrease in body weight was evident in the high-dose groups (36% decrease in males, 24% decrease in females). Blood hemoglobin levels were reduced in both sexes in the high-dose group and in mid-dose females. Biochemical changes in the high-dose group consisted of increases in albumin, cholesterol, urea, and GGT activity. Liver and kidney weights were increased in a dose-dependent manner. Histopathologic examination revealed hypertrophic hepatocytes (containing a basophilic granular substance) and vacuolation of hepatocytes in periportal areas. Coagulative necrosis and oval cell proliferation were occasionally seen. Atrophy of the epithelial lining of the proximal convoluted tubules in the renal cortex was present in some male rats. A NOAEL for PBO could not be established in this study owing to the presence of liver and kidney effects even at the “low” dose of 6000 ppm. Marked clinical signs of subacute toxicity were seen in the dams of a range-finding study conducted to select doses for a developmental toxicology study (Chun and NeeperBradley, 1992). Pregnant female rats (15 animals/dose level) were gavaged on gestational days 6–15 with PBO at levels of 250, 500, 1000, 2000, or 4000 mg/kg body weight/day. Signs of general stress, such as urogenital wetness and periocular encrustation, were evident in many animals during the first three days of the study at dose levels of

Chapter | 99  The Safety Assessment of Piperonyl Butoxide

at least 500 mg/kg body weight/day. At levels of 2000 mg/kg body weight/day, more severe clinical signs such as ataxia, twitching, prostration, dyspnea, gasping, and lacrimation were noted. Ulceration of the lining of the glandular region of the stomach as well as hemorrhage and sloughing of the lining of the nonglandular region were noted at necropsy. In the only subchronic study conducted by the inhalation route, Newton (1992) exposed CD rats (15 animals/sex/dose level) for 6 h/day, 5 days/week, for 90 days in whole body exposure chambers. PBO was aerosolized to achieve exposure concentrations of 15, 74, 155, and 512 mg PBO/m3 (MMAD of the aerosol was 1.7 m). Neither body weight gain nor food intake was affected by exposure. In the high-dose group, serum alanine transaminase, aspartate transaminase, and glucose levels were decreased, whereas BUN, total protein, and albumin levels were increased. However, not all of these effects were statistically significant, and there was no clear dose–response relationship. Both absolute and relative liver and kidney weights were elevated in the high-dose group. Minimal to slight irritation of the larynx was observed upon necropsy in all treatment groups. Inflammation, congestion, edema, and debris in the lumen were noted as well. Squamous metaplasia of the laryngeal epithelium was noted in all groups but was most marked in both sexes at the highest dose level. Important questions arise from the observation of metaplasia: Does the effect represent true toxicity that is likely to progress to neoplasia upon chronic exposure or does it represent a limited adaptive response to mild irritation? l Does the observation of squamous metaplasia in the rat laryngeal epithelium have relevance for humans? l

A recent comprehensive review of the literature showed that laryngeal metaplasia can be produced by a wide range of chemically dissimilar substances, and even by “nonchemical” means such as irritation by aerosols and particles, and dehydration by alcohols or low humidity air (Osimitz et al., 2009). Other factors that indicate that this response is adaptive and self-limiting include: The well-differentiated character of laryngeal squamous metaplasia l The reversibility of incidence and severity of it during recovery periods demonstrated with several chemicals l The lack of significant clinical observations associated with the effect l The lack of progression of the lesion over time l

Moreover, by virtue of their anatomy, the rat is more sensitive to irritation of the tissue in the larynx than is the human. Gopinath et al. (1987) emphasized that “it would appear that the rodent larynx is particularly sensitive to aerosol damage.” Thus the same dosage of an inhaled chemical delivered to rodent and human larynx would be more likely to cause histopathological alterations in the rat larynx than the human larynx. Taken together, the squamous

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metaplasia of the rodent larynx should not be used as a toxicologic endpoint for quantitative risk assessment.

99.3.2.4  Rabbits Sarles et al. (1949) conducted a subacute oral toxicity experiment in rabbits. A 5% PBO emulsion was fed once weekly in the diet to three rabbits over a 3-week period. The dosage used varied between 1.0 and 4.0 ml/kg body weight/ week. There was neither mortality nor clinical signs of toxicity. The rabbit that received the highest dosage was sacrificed 1 week after the last treatment, but no lesions were detected at postmortem examination. A 21-day subchronic dermal toxicity study was conducted in rabbits (five animals/sex/dose level) in which 100, 300, or 1000 mg PBO/kg body weight was applied topically once a day, 5 days a week, for 3 consecutive weeks (Goldenthal, 1992). Treatment-related effects were limited to minor skin changes at the application site. Dermal irritation was present in all treatment groups (although to a lesser extent and incidence at the 100 mg/kg body weight/day dose level). Dermal lesions consisted of very slight erythema and edema. This irritation usually appeared by day 5 and persisted for the remainder of the study. Desquamation and fissuring of the skin appeared in the 300 and 1000 mg/kg body weight/day groups. Moderate acanthosis, hyperkeratosis, and chronic inflammation of the epidermis were present. The severity of these lesions increased with increasing dosage. Body weights were comparable with those in the control group, and food intake was only slightly lower in treated animals. No treatmentrelated changes were seen in hematology and clinical chemistry, and no signs of systemic toxicity were present at any dosage level. The NOAEL is 100 mg/kg body weight/day for local effects, whereas the NOAEL for systemic toxicity is 1000 mg/kg body weight/day.

99.3.2.5  Other Species A 4-week oral toxicity study was conducted by Sarles and Vandergrift (1952) in which two African Green monkeys were fed PBO by capsule, 6 days a week (for 4 weeks), at a dosage level of 0.03 or 0.1 ml/kg body weight/day (one monkey at each dosage level). No gross pathological lesions were evident in the treated monkeys’ livers. Upon histopathologic examination of the liver, the monkey on the higher dose level showed evidence of minimal dystrophy and dysplasia, occasional acidophilic and hyaline-necrosis cells, as well as hydropic swelling.

99.3.3  Reproductive Toxicity A summary of the results of the reproductive and developmental toxicity studies discussed below is presented in Tables 99.3 and 99.4.

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Table 99.3  Summary of Results from Reproductive Studies with PBO Species Route Dose

Study type

Mice

Three-generation Not established

Rat

Dog

NOAEL

Comments

Reference

Pup weights reduced at all dose levels.

Tanaka et al. (1992)

Diet

1000–8000 ppm (268–1583 mg/kg body weight/day)

Diet

1500–6000 ppm Two-generation (400–1250 mg/kg body weight/day)

Not established

Excessive dose levels resulted in decreased pup weights in all treated animals.

Tanaka (1992)

Diet

1500-6000 ppm – males only

Postnatal

3000 ppm

High-dose animals showed increase in motor activity parameters at 11 weeks.

Tanaka (1993)

Diet

100–900 ppm (30–250 mg/kg/ body weight/day)

Two-generation

100 ppm for neurobehavioral effects

Surface righting delayed at PND 7 (300 Tanaka (2003) and 900 ppm). Olfactory orientation reduced at PND 14 (300 and 900 ppm). Total distance in a motor activity monitor increased at 9 weeks (300 and 900 ppm). Average distance and average speed in a motor activity monitor increased at 9 weeks (900 ppm).

Diet

100–25,000 ppm (8–2000 mg/kg body weight/day)

Three-generation 1000 ppm (8 mg/kg body weight/day)

Diet

300–5000 ppm (24–400 mg/kg body weight/day)

Two-generation

Parental toxicity/ Body weights of pups born to dams pup development: treated at the highest dose level were 1000 ppm (80 mg/ reduced in the early postnatal period. kg body weight/day) Reproductive toxicity: 5000 ppm (400 mg/ kg body weight/day)

Diet

30–500 mg/kg body weight/day

2-yr chronic

Not applicable

Increases in ovarian weight observed in Butler et al. some females at highest dose level. (1998)

Diet

500–3000 ppm (12.5–75 mg/kg body weight/day)

Range-finding

Not applicable

Increased absolute and relative weights Goldenthal of testis and epididymis noted. No (1993a) microscopic abnormalities observed in the testis.

99.3.3.1  Dogs In an 8-week dietary range-finding toxicity study in dogs, PBO was fed daily in the diet to beagles at dose rates of 500, 1000, 2000, or 3000 ppm (approximately equivalent to 12.5, 25, 50, or 75 mg/kg body weight/day, respectively) (Goldenthal, 1993a). There was an increase in the absolute and relative weights of the testes and epididymides. No microscopic abnormalities were noted in the testes. Spermatozoa were being produced. Other details of the study are presented in Section 99.3.2.1.

99.3.3.2  Mice A three-generation, one litter per generation, reproductive study was conducted in CD-1 mice by Tanaka et al. (1992).

Very high maternal toxicity at two highest dose levels resulting in marked reductions in the incidence of pregnancies, numbers of litters per dam, general health of the offspring, and average weanling weights of pups.

Sarles and Vandergrift (1952)

Robinson et al. (1986)

Ten animals of either sex were fed diets containing 1000, 2000, 4000, and 8000 ppm PBO (purity not specified). These doses are equivalent to 268, 506, 936, and 1583 mg PBO/kg body weight/day as averaged over F0 and F1 generations from preconception to lactation. Food intake was reduced in the F0 generations at the 8000 ppm dose, indicative of some toxicity, except during the mating period, and was also reduced during the lactation period in the F1 generation, also at 8000 ppm. The 4000 ppm treatment groups of both the F0 and F1 generations also had a reduction in food intake during the lactation period. Mean F1 litter weight was significantly decreased (38%) at 8000 ppm and reduced by 18% at the 4000 ppm treatment level. However, litter size remained unchanged at all levels. Pups born in the 8000-ppm-treated group had a lower survival index at postnatal day 21 (63% vs. 91% for males of control group; 79%

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Table 99.4  Summary of Results from Developmental Studies with PBO Species Route

Dose

NOAEL

Comments

Mice

Gavage

1065–1800 mg/kg body weight/day

Not established

Total resorption rates significantly Tanaka et al. increased in mid- and high-dose (1994) groups. Significant decrease in body weights of male and female fetuses, appearing to be dose-dependent.

Rat

Gavage

300 or 1000 mg/kg body weight/day

Maternal body weights were Maternal toxicity: Not established Developmental toxicity: 1000 mg/kg reduced at both dose levels tested. body weight/day

Gavage

62.5–500 mg/kg body Maternal and developmental No signs of either maternal or weight/day toxicity: 500 mg/kg body weight/day embryo-fetotoxicity.

Gavage

200–1000 mg/kg body weight/day

Maternal toxicity: 200 mg/kg body weight/day

Reference

Kennedy et al. (1977) Khera et al. (1979)

Gestational body weights and body Chun and weight gains were reduced in the Neeper-Bradley 500- and 1000-mg/kg body weight/ (1991) day groups.

Developmental toxicity: 1000 mg/kg body weight/day Gavage

Rabbit

Gavage

630-1800 mg/kg body Maternal toxicity: 630 mg/kg body weight/day weight/day

50–200 mg/kg body weight/day

Decreased maternal weight gain in 1065- and 1800-mg/kg body weight/ day groups.

Tanaka (1995)

Developmental toxicity: 630 mg/kg body weight/day

Decreased fetal body weight, increased external limb deformities in 1065- and 1800-mg/kg body weight/ day groups.

Maternal toxicity: 50 mg/kg body weight/day

Maternal toxicity was evident at Leng et al. (1986) 100 and 200 mg/kg body weight/day manifested by decreased defecation and a dose-dependent weight loss during the treatment period.

Developmental toxicity: 200 mg/kg body weight/day

vs. 89% for females of the control group). Pup weights in the F1 generation were decreased for all dosage groups, but there was no dose-related response at low- or mid-dose levels. No treatment-related effects were noted in neurobehavioral tests conducted in the F1 animals. The mean F2 litter size was significantly decreased at the 4000 and 8000 ppm treatment levels. Mean F2 treatment weights were decreased in all treatment groups. Pups in the 8000-ppmtreated group of the F2 generation had a lower survival index than controls at day 21 postnatal (59% in males and 79% in females vs. 100% in male and female control groups). Pup weights in the F2 generation were decreased at dosage levels of 2000 ppm and above on postnatal days 4, 7, 14, and 21. Pup weights were reduced on days 4 and 7 at the 1000 ppm dosage level. Changes in neurobehavioral parameters such as surface righting and cliff avoidance at postnatal day 7 were noted in the F2 pups. However, as the author pointed out, no clear dose–response relationship was evident in these neurobehavioral tests. Because pup weights

were reduced at all doses tested, a NOAEL for reproductive effects could not be established from this study.

99.3.3.3  Rats Sarles and Vandergrift (1952) report a study in which groups of 12 male and 12 female rats per dose level were fed diets containing 100, 1000, 10,000, or 25,000 ppm (approximately equivalent to 8, 80, 800, or 2000 mg/kg body weight/ day, respectively) of PBO (technical grade, 80% purity), for three generations. None of the female rats at the highest dose level were fertile and there were marked reductions in the incidence of pregnancies, numbers of litters per dam, general health of the offspring, and average weanling weights of pups born to dams treated at 10,000 ppm. These findings are clearly a result of the high maternal toxicity, especially at 10,000 and 25,000 ppm. No adverse effect on reproduction was observed in three generations of progeny in the 100 and 1000 ppm groups (NOAEL  1000 ppm).

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A two-generation reproduction study was conducted in rats with PBO by Robinson et al. (1986). Groups of 26 male and 26 female Sprague-Dawley rats were utilized and adults of the F0 and F1 generations were treated at dose levels of 300, 1000, or 5000 ppm (approximately equivalent to 24, 80, or 400 mg/kg body weight/day, respectively) in the diet. Animals were treated for 83–85 days prior to placement for mating, and treatment continued throughout mating, pregnancy, and lactation. The only consistent finding throughout the study period was a lower body weight gain at the highest dosage level. This tendency was partially reversed during the lactation period, when females at this dose level showed higher weight gains when compared with control rats. For both F1 and F2 generation pups, the viability, survival, and lactation indices were unaffected by treatment. There were no treatment-related abnormal findings for the pups, and weanlings did not reveal any treatment-related adverse effects. Body weights of pups born to dams treated at the highest dose level were reduced in the early postnatal period. The NOAEL for parental toxicity and pup development was thus 1000 ppm PBO in the diet. The NOAEL for reproductive toxicity was set at 5000 ppm PBO in the diet. In a 2-year chronic oral toxicity study in Sprague-Dawley rats, animals received dietary administration of 15, 30, 100, or 500 mg PBO/kg body weight/day (89% purity) (Butler et al., 1998). Only changes in the reproductive system are discussed here. Full details of the study are presented in Section 99.3.4.3. Increases in ovarian weights were observed among some females receiving 500 mg/kg body weight/day, although no histopathologic changes were noted. Atrophy of the testes was seen histologically in all male groups and when bilateral atrophy was considered alone, there was an increased incidence in the intermediate and high-dose groups with a corresponding reduction in the incidence of unilateral atrophy. However, the finding is unlikely to be related to treatment because the dose–response relationship was unclear and the atrophy was not accompanied by changes in the seminiferous tubules or sperm production. Moreover, there were no statistically significant increases or decreases in testes weight when expressed as either absolute weight or relative-to-brain weight.

99.3.4  Developmental Toxicity 99.3.4.1  Mice Tanaka et al. (1994) reported a study conducted in CD-1 mice in which PBO was administered by gavage on day 9 of gestation to groups of 20 animals at doses of 1065, 1385, or 1800 mg/kg body weight (95% purity; PBO dissolved in olive oil). No abnormal behavior or mortality patterns were observed in dams. Three abortions occurred in the mid- and high-dose groups. Four litters were resorbed in the two higher-dosage groups, but maternal body weights

were comparable between all groups. Total resorption rates were significantly increased in the mid-dose (26%) and the high-dose groups (32%) when compared with the control value (6%). The number of viable fetuses per dam was comparable between all dosage groups. There was a significant decrease in body weights of male and female fetuses derived from treated dams, which did appear to be dose-dependent. Certain external malformations such as exencephaly, cranioschisis, open eyelids, omphalocele, kinky tail, and talipes varus were observed in all groups (including controls) and oligodactyly was recorded in the forelimbs of some fetuses derived from treated dams. The incidence of this latter defect was 6% in those fetuses derived from the highest dosage group. The authors concluded that a single high dose of PBO (1065 mg/kg body weight or above), when given orally to pregnant mice on day 9 of gestation, could cause embryo-fetal toxicity with associated oligodactyly of the forelimbs. The high dose levels of this study make it difficult to interpret the significance of this finding.

99.3.4.2  Rats Kennedy et al. (1977) carried out a developmental toxicity study with PBO in pregnant rats. Twenty female animals per dose level were gavaged with PBO in corn oil at 300 or 1000 mg PBO/kg body weight/day. Other than a decline in body weight gain in both treated groups (especially in the later stages of gestation), no other treatment-related signs of toxicity occurred. The reproductive parameters of the dams were not significantly affected by treatment. One female from each treatment group resorbed most or all of her litter. The fetuses derived from each treatment group exhibited no internal or external skeletal malformations that could be related to treatment. The observation that maternal body weights were reduced at both doses tested meant that a NOAEL was not established for maternal toxicity. Because no developmental effects were seen at any dose including the highest dose tested, the NOAEL for embryo-fetal toxicity for this study was 1000 mg/kg body weight/day. Pregnant female Wistar rats (17–20 per dosage group) were dosed with PBO levels of 62.5, 125, 250, or 500 mg/kg body weight/day from day 6 to day 15 of gestation in a study conducted by Khera et al. (1979). The types and incidences of anomalies in fetuses derived from treated dams were comparable with those of the control group and it was concluded that doses as high as 500 mg/kg body weight/day produced no signs of either maternal or embryo-fetal toxicity. A developmental toxicity study by Chun and Neeper-Bradley (1991) conducted in accord with U.S. EPA Guidelines investigated PBO in Sprague-Dawley rats. Timed pregnant rats were administered PBO (90.78% purity) by gavage on gestation days 6–15. The dosage

Chapter | 99  The Safety Assessment of Piperonyl Butoxide

l­evels were 200, 500, and 1000 mg/kg body weight/day and 25 animals were included in each group. The pregnancy rate was equivalent among groups and ranged from 88% to 96%. No females aborted, delivered early, or were removed from the study. Gestational body weights and body weight gains were reduced in the 500 and 1000 mg/kg body weight/day groups, as was food intake for the first 7 days, indicating that a sufficiently high dose was achieved. Treatment had no effect on gestational parameters including resorption, pre- and postimplantation losses, percentage of live fetuses, and sex ratios, nor did it affect the fetal body weights or the incidence of fetal malformations. However, two common skeletal variations (i.e., nonossification of centrum of vertebrae 5 or 6) had a higher incidence in the two highest dosage groups. These findings were not considered treatment-related, as adjacent vertebrae did not have delayed ossification. The NOAEL for maternal toxicity in the rat was 200 mg/kg body weight/day and the NOAEL for developmental toxicity was at least 1000 mg/kg body weight/day. Tanaka et al. (1995) reported a development toxi­city study in pregnant female Charles River (Crj:CDS) rats dosed with PBO (purity 95%) at levels of 0, 630, 1065, and 1800 mg/kg body weight on days 11–12 of gestation. The most significant treatment-related changes included an increased resorption at the high dose and various limb deformities (oligodactyly, syndactyly, and polydactyly) in the 1065- and 1800-mg/kg body weight dose groups. Two of 10 dams and six of eight dams had anomalous fetuses in the 1065- and 1800-mg/kg body weight dose groups, respectively. No anomalous fetuses were seen at the 630-mg/kg body weight level. These results must be interpreted carefully. Although dosing had no effect on mortality of the treated dams, a large and statistically significant decrease in maternal body weight gain (days 11–20) was seen at 1065 and 1800 mg/kg body weight (23.7% and 36.1%, respectively). Correspondingly, although treatment had no effect on the average litter size, the average fetal body weight of males was decreased at 1065 mg/ kg body weight (4.8%) and for both male and females at 1800 mg/kg body weight (14.5% and 13.5% respectively). It is uncertain what role these significant weight changes have on the development of the limb abnormalities, but the potential significance of those observations is limited given the possible confounding effects of the hightreatment dose levels on the dams.

99.3.4.3  Rabbits New Zealand White female rabbits were gavaged with PBO (purity 100%) in corn oil at levels of 50, 100, or 200 mg/kg body weight between day 7 and day 19 of pregnancy (Leng et al., 1986). Caesarian sections were performed on day 29 of gestation. Maternal toxicity was evident at 100 and 200 mg/kg body weight/day manifested by decreased

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defecation and a dose-dependent weight loss during the treatment period (these weight losses were recovered posttreatment). Common developmental defects, including an increase in the number of full ribs and the presence of more than 27 presacral vertebrae, were recorded in all dose groups. However, no dose–response relationship was apparent. The number of litters in the treated groups with these observations was not increased when compared with control values. The NOAEL for maternal toxicity was 50 mg/ kg body weight/day, whereas the NOAEL for developmental toxicity was 200 mg/kg body weight/day.

99.3.4.4  Developmental Neurotoxicity A three-generation reproductive study that included neuro­ behavioral endpoints was carried out in CD-1 mice by Tanaka et al. (1992). Ten animals of either sex were fed diets containing 1000, 2000, 4000, and 8000 ppm PBO (purity not specified). These doses are equivalent to 268, 506, 936, and 1583 mg PBO/kg body weight/day as averaged over F0 and F1 generations from preconception to lactation. No treatment-related effects were noted in neurobehavioral tests conducted on the F1 or F2 animals. Tanaka (1993) further studied the developmental neurotoxicity of PBO (purity not specified) in CD-1 mice in a subsequent two-generation study. Ten mice of either sex per group received diets containing 1500, 3000, or 6000 ppm (approximately equivalent to 400, 700, and 1250 mg/kg/ body weight/day) PBO during a 4-week period prior to mating (F0), during gestation and through the time that the F1 generation was 9 weeks old. Maternal body weights and food consumption were not reported. The open field test demonstrated a statistically significant decrease in ambulating in F0 male mice only at the highest dose (6000 ppm). Because of the excessive dosage levels incorporated in this study, pup body weights were reduced at birth at all dosed animals. By postnatal day (PND) 21, the mean pup body weights of mediumand high-dose pups were decreased by 9.4% and 41%, respectively. The survival index for pups at postnatal day 21 was 79.2% (controls), 92.9% (low-dose group), 80.0% (mid-dose group), and 51.7% (high-dose group). There were no significant differences in the behavioral tests during the lactation period, except for a reduction in olfactory orientation in mid- and high-dose group animals (41.2% and 39.2%, respectively). Other than sporadic nondose-dependent changes, the open field test and multiple water T-maze tests were not significantly altered by PBO. In a postnatal developmental neurotoxicity study, Tanaka (1993) administered PBO to male mice only from 5 to 12 weeks of age at levels of 1500, 3000, and 6000 ppm in the diet. PBO had no effect on food consumption; body weights were not reported. The high-dose animals exhibited motor activity effects at 11 weeks of age, including an increase in number of movements, movement time, total distance, average speed, and number of turnings.

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Tanaka (2003) conducted another two-generation developmental neurotoxicity study in CD-1 mice using dietary dose levels of 100, 300, and 900 ppm from 5 weeks of age in the F0 through 9 weeks of age in F1. No effects were noted in the number of offspring or the survival of offspring through lactation. Treatment had no effect on the body weights of mice after weaning. The following effects were noted in males only:

Test Guidelines OPPTS 870.6300 Developmental Neuro­ tox­icity Study (U.S. EPA, 1998). Moreover, a summary of developmental studies in which neurobehavior was evaluated indicated that out of 191 such studies, only three were conducted in the mouse (Ulbrich and Palmer, 1996).

Surface righting was delayed at PND 7 (300 and 900 ppm). l Olfactory orientation was reduced at PND 14 (300 and 900 ppm). l Total distance in a motor activity monitor was increased at 9 weeks (300 and 900 ppm). l Average distance and average speed in a motor activity monitor was increased at 9 weeks (900 ppm).

Numerous long-term toxicity and oncogenicity studies have been undertaken on PBO over the past 50 years in various species. As evidenced in these studies, the primary target organ is the liver. The results of the studies discussed below are summarized in Table 99.5.

These effects were noted at levels significantly lower than reported in previous studies. Tanaka (1992) previously showed that PBO depressed olfactory orientation in the F1-generation mice in a two-generation toxicity study (1500–6000 ppm in the diet). Tanaka et al. (1992) reported that PBO depressed olfactory orientation in the F1-generation mice and effects on several other parameters in F2generation mice were observed (1000–8000 ppm in the diet). It is not known what accounts for the seemingly high sensitivity of the male mice in this Tanaka (2003) study. Regarding the Tanaka studies, a low incidence of statistically significant measures in a few behavioral measurements should not be taken per se as sufficient evidence of developmental neurotoxicity. Developmental neurotoxicity studies evaluate a wide range of behavioral and histopathological factors. An expert panel has recently evaluated developmental neurotoxicology endpoints (Tyl et al., 2008). This panel has recognized that behavioral endpoints, in particular given the intrinsic variability of age-related development in very young animals, may be highly variable. Rather than drawing conclusions from single “statistical significances,” it is necessary to look for consistent patterns of behavioral effect in “functional domains” and using statistics appropriate to multiple measures over time. Meanwhile, behavioral endpoints must also be taken in the context of systemic toxicity. Test rodents, like their human counterparts, suffer from clinical and behavioral symptoms due to systemic illness (Gerber and O’Shaughnessy, 1986, reference 18). Moreover, developmental neurotoxicity observations in the mice are difficult to evaluate given the paucity of such data in mice. Although rabbits and rats, and, to a lesser extent mice, are the animal species primarily used in routine developmental toxicity, testing the assessment of neurobehavioral toxicity in the offspring is ­evaluated almost exclusively in rats. Rats are the preferred test species in the OECD Test No. 426: Developmental Neurotoxicity Study (OECD, 2001) and the U.S. EPA Health Effects

PBO was administered to dogs in capsule form for a 1-year chronic dietary toxicity study (Sarles and Vandergrift, 1952). Groups of four dogs each were treated at dose levels corresponding to 3, 32, 160, or 320 mg/kg body weight/day. The dosage was adjusted in accordance with any alteration in body weight to maintain the same dose in mg/kg body weight, except for one individual animal per dosage group, which received a constant absolute dose throughout the trial. All dogs belonging to the two highest dosage groups lost weight; however, meaningful comparisons between the lower dose groups and control animals were not possible owing to large variations in body weight gains and the small number of animals involved. All dogs at the highest dosage level died. However, no toxic reaction was seen at 3 mg/kg body weight/day. Red blood cell (RBC) and white blood cell (WBC) counts were unchanged at all dose levels. There was a dose-dependent increase in liver, kidney, and adrenal weights. Microscopic changes were quite similar to those in long-term toxicity studies conducted in rats, with the liver again being the major target organ for toxicity. Hydropic swelling was evident in hepatocytes in the mid-dose group, with hepatic dystrophy and dysplasia becoming more obvious at the two highest dosage levels. The NOAEL for this study was 32 mg/kg body weight/day. A 1-year chronic dietary toxicity study following USPEA Guidelines was conducted in the beagle dog in which groups of four males and four females were fed PBO for 1 year at doses of 100, 600, or 2000 ppm (approximately equivalent to 2.5, 15, or 50 mg/kg body weight/day, respectively) in the diet (Goldenthal, 1993b). All animals survived to study termination. A reduction in body weight gain and food intake was evident in the 2000 ppm group. Physical examinations were otherwise normal throughout the test period. Biochemical analysis showed increases in serum alkaline phosphatase levels at 6 and 12 months in the highest dosage group. Female beagles showed a decrease in serum cholesterol at the 2000-ppm dosage level. Increased liver and gall bladder weights, with mild

l

99.3.5  Chronic Toxicity/Oncogenicity

99.3.5.1  Dogs

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Table 99.5  Summary of Results of Chronic Toxicity/Oncogenicity Studies with PBO Species Route

Dose

Duration

NOAEL

Comments

Reference

Dog

Oral

3–320 mg/kg body weight/ day

1 year

32 mg/kg body weight/day

Increased liver and kidney weights, hepatic dystrophy and dysplasia

Sarles and Vandergrift (1952)

Diet

100–2000 ppm (2.5–50 mg/ 1 year kg body weight/day)

Diet

300 or 1112 ppm (45 or 167 mg/kg body weight/day)

69 weeks

Diet

45 or 133 mg/kg body weight/day

18 months 45 mg/kg body weight/day

No signs of toxicity

Bond et al. (1973)

Diet

1036–2804 ppm (148– 298 mg/kg body weight/day)

112 weeks Not established

Decreased body weight in both sexes, hepatic nodular hyperplasia in males

U.S. National Cancer Institute (1979)

Diet

6000–12,000 ppm (960– 1 year 1920 mg/kg body weight/day)

Not established

Hepatic adenomas and hepatocarcinomas

Takahashi et al. (1994b)

Diet

30–300 mg/kg body weight/ day

30 mg/kg body weight/day

Increased liver weight, benign hepatic adenomas

Butler et al. (1998)

Diet

2 year 100–25,000 ppm (5– 1250 mg/kg body weight/day)

100 ppm (5 mg/kg body weight/day)

No significant increase in tumor Sarles and incidence; severe liver damage, Vandergrift (1952) increased incidence of liver “hyperdysplastic” nodules

Diet

90 mg/kg body weight/day

2 year

Not applicable

Decreased body weight

Diet

5000–10,000 ppm (~250– 500 mg/kg body weight/day)

107 weeks Not established

Dose-related increase in Cardy et al. (1979) hepatocytomegaly, dose-related increase in lymphomas in female rats, but incidence in controls was also high

Diet

5000–10,000 ppm (250– 500 mg/kg body weight/day)

2 year

Not established

No significant increase in tumor Maekawa et al. incidence; dose-related increase (1985) in ileocecal ulcers

Diet

6000–24,000 ppm (526– 95–96 2187 mg/kg body weight/day) weeks

Not established

Hepatic adenomas and hepatocarcinoma at midand high-dose levels, cecal hemorrhaging, severe general and hepatic toxicity

Takahashi et al. (1994a)

Diet

15–500 mg/kg body weight/ day

2 year

30 mg/kg body weight/day

Increased liver and kidney weights, centrilobular hepatocyte hypertrophy, focal hyperplasia

Butler et al. (1998)

Diet

2.0 ml/day (1000 ppm)

1 year

Not applicable

Slight hepatic dystrophy and dysplasia

Sarles and Vandergrift (1952)

Mouse

Rat

Goat

78 weeks

600 ppm Increased liver and gall bladder (15 mg/kg body weights, hypertrophy of weight/day) hepatocytes Not applicable

hypertrophy of hepatocytes, were also recorded at this highest dosage level. A small increase in thyroid gland and parathyroid gland weights was also noted. However, no microscopic abnormalities were detected in the thyroid gland. No treatment-related histopathologic changes were seen on the study. Based on the changes seen in the liver, the NOAEL for this study was 600 ppm.

Goldenthal (1993b)

No significant increase in tumor Innes et al. (1969) incidence

Hunter et al. (1977)

99.3.5.2  Mice Innes and co-workers (1969) studied the oncogenicity of PBO in mice by administering the maximal tolerated dose. Animals (18/sex/strain) from two hybrid stocks (C57BL/ 6  C3H/Anf or C57BL/6  AKR) were gavaged with 100 mg undiluted PBO/kg body weight or 464 mg PBO/kg

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body weight in solvent vehicle from 7 to 28 days of age. Thereafter, they received 300 ppm undiluted PBO or 1112 ppm PBO in solvent vehicle (approximately equivalent to 45 and 167 mg/kg body weight/day, respectively) in the diet for 69 weeks. These researchers found no significant increase in tumor incidence as a result of PBO treatment. Bond et al. (1973) reported no adverse effects following dosing of mice with 45 or 133 mg/kg body weight/day PBO in the diet for 18 months. Few details are available for this early study, however. The U.S. National Cancer Institute (1979) conducted a mouse carcinogenicity study in which male and female B6C3F1 animals (50/sex/dose level) were initially dosed with 2500 or 5000 ppm of PBO in the diet. Toxicity appearing in both these groups resulted in a reduction in the doses to 500 and 2000 ppm, respectively, after 30 weeks of dosing. The time-weighted average doses in the diets were approximately 1036 and 2804 ppm (approximately equivalent to 148 and 298 mg/kg body weight/day, respectively). Dose-dependent decreases in body weight and body weight gains were observed in both dose groups and both sexes. Nodular hyperplasia of the liver was slightly elevated in males. Although tumors were observed in the liver and lacrimal gland, the incidence was not statistically significant. Thus, the authors concluded that PBO was not oncogenic. Takahashi et al. (1994b) reported the results of a 1-year chronic toxicity study conducted in CD-1 mice using dietary doses of 6000 and 12,000 ppm PBO (equivalent to 960 and 1920 mg/kg body weight/day, respectively). Animals were allocated to three groups consisting of 52, 53, and 100 animals for dose levels of 0, 6000, and 12,000 ppm, respectively. Significant depressions in body weight and body weight gain were noted for low and high doses of PBO. Only 81% of the high-dose animals survived the 12-month study period compared with 98% and 94% of the animals in the low-dose and control groups, respectively. Hepatic adenomas and hepatocarcinomas were observed in both treatment groups as were hemangiosarcomas and hemangioendothelial sarcomas. However, the doses used in this study are clearly in excess of internationally accepted maximum tolerated dose (MTD) criteria and thus they are of questionable relevance in determining the hazard and risk for humans of PBO exposure. Butler et al. (1998) reported a U.S. EPA Guideline study in which groups of 60 male and 60 female CD-1 mice were administered PBO in the diet at doses of 0 (two separate control groups), 30, 100, or 300 mg/kg body weight/day for at least 78 weeks. No treatment-related clinical signs of toxicity or changes in food consumption or clinical chemistry were observed. The mean absolute body weight and mean body weight gains were generally slightly decreased throughout the study at the high dose in both males and females, indicating that the MTD was reached. A dose-related increase in the mean absolute and relative liver weights was seen in the mid- and high-dose groups of both sexes. The mean

absolute and relative liver weights of the low-dose group of male mice were also slightly increased. Both males and females clearly showed an increased incidence of benign hepatic nodules diagnosed as adenomas. The further characterization of the adenomas showed that the increased burden of lesions was due to the increased incidence of eosinophilic adenomas, similar to the lesions induced by a range of enzyme inducers in the mouse (Butler, 1996). There was no increase of either basophilic adenomas or hepatocarcinomas. The NOAEL in this study for nononcogenic effects was 30 mg/kg body weight/day.

99.3.5.3  Rats In an early study, Sarles and Vandergrift (1952) fed Wistar rats with diets containing from 100 to 25,000 ppm (approximately equivalent to 5–1250 mg/kg body weight/day) PBO for 2 years. Twelve males and 12 females were used at each dose level. The entire high-dose group died by week 68 and showed severe liver damage upon necropsy. An increased incidence of “hyperdysplastic” hepatic nodules, characterized by the authors as the appearance of larger cells and increased polyploidy, was seen in the treated groups. Dystrophy and dysplasias were also observed in the livers from animals fed 1000 ppm or greater PBO. The authors concluded that there was no evidence of carcino­ genicity; the 100 ppm dose level was considered nontoxic. Because PBO is most often used as a synergist with pyrethrins, a 2-year dietary study was conducted in SpragueDawley rats using a mixture of pyrethrins (53.1% purity) and PBO (95% purity) (Hunter et al., 1977). Forty-five males and 45 females were fed diets containing 400 ppm pyrethrins plus 2000 ppm PBO. The average daily doses received by the animals over the study period were 16  79 mg/kg body weight/ day (pyrethrins  PBO) for males and 20  101 mg/kg body weight/day (pyrethrins  PBO) for females. Body weights were depressed in the females during the first 78 weeks of treatment and among males during the first 26 weeks of treatment. No other treatment-related effects were noted and no treatment-related change in tumor incidence was seen. The U.S. National Cancer Institute conducted a 2-year cancer bioassay in Fisher 344 rats (Cardy et al., 1979). Fifty male and 50 female animals per dose level were allocated to low- and high-dose groups which received PBO (88.4% purity) in the diet at 5000 or 10,000 ppm (approximately equivalent to 250 or 500 mg/kg body weight/day, respectively) for 107 weeks. A dose-dependent decrease in the mean body weights of treated groups was noted. Other than increased hepatocytomegaly, no dose-related increases in the incidence of tumors or other microscopic findings were observed in the liver. The hepatocytomegaly consisted of foci of enlarged hepatocytes, often associated with large, vesicular nuclei and numerous cytoplasmic vacuoles, giving the cytoplasm a “ground glass” appearance.

Chapter | 99  The Safety Assessment of Piperonyl Butoxide

Distortion of lobular architecture in these foci was minimal, and trabeculae were continuous with adjacent normal hepatocytes. These lesions appear similar to those described by Squire and Levitt (1975) as “eosinophilic foci,” “ground glass foci,” or “clear cell foci.” Although a dose-dependent increase in lymphomas was noted in females, the incidence of lymphomas, leu­kemias, and reticuloses observed was not significantly different from the historical rates from the laboratory. Thus, this study showed that, under the conditions of the bioassay, PBO was not carcinogenic in Fischer 344 rats. The oncogenicity of PBO was also studied in F344/ DuCrj rats by Maekawa et al. (1985). Animals (50 sex/ dose level) were fed a dietary level of 5000 or 10,000 ppm (approximately equivalent to 250 and 500 mg/kg body weight/day, respectively) for 2 years but no significant dose-related increase in the incidence of any tumor was found. A dose-related incidence of ileocecal ulcers, however, was found in animals of both sexes. Takahashi et al. (1994a) conducted a 2-year chronic toxicity study in the rat at dose levels up to 24,000 ppm (1000 times the maximum level approved in raw cereals in Japan). Fischer F344/DuCrj rats (30–33 per group) received a diet containing PBO at 6,000, 12,000, or 24,000 ppm (equivalent to 526, 1052, or 2187 mg/kg body weight/day, respectively) for 95–96 weeks. Beginning at about 40 weeks, 10 rats in the 12,000 ppm male group died due to cecal hemorrhages. By the end of the study, gastrointestinal hemorrhage occurred at all dose levels. Organ weights (with the exception of the liver) were reduced in all animals in the high-dose group. “Probable essential thrombocytopenia” was present in all treated male groups. Body weight gains relative to controls were reduced in all treated groups of both sexes and reached approximately 50% in the highdose group. A dose-dependent increase in hepatocellular hyperplasia (seen as liver nodules) was reported. Although the nomenclature is different, these lesions are much like those described by Sarles and Vandergrift (1952) at toxic doses of PBO. Takahashi also reported hepatocellular adenomas and carcinomas in the mid- and high-dose groups. It is important to note that the study was not intended to be a carcinogenicity study and the procedures for collecting and examining tissues were not performed according to current U.S. EPA/OECD standards. Thus, not all tissues were taken or prepared for histological examination. Because of the high-dose levels used and resulting tox­icity, it is difficult to interpret the carcinogenicity findings and their relevance for hazard assessment. Moreover, several investigators have reported that hepatotoxicity, and the resulting regenerative hyperplasia, can contribute to the formation of liver tumors by nongenotoxic mechanisms (Kociba et al., 1978; McClean et al., 1990; Mutai et al., 1990; Tatematsu et al., 1990; Van Miller et al., 1977). This and other mechanistic aspects of this oncogenicity response are discussed in Section 99.3.5.

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Butler et al. (1998) reported the results of a U.S. EPA Guideline dietary study in the Sprague-Dawley rat in which groups of 60 animals of each sex were administered 15, 30, 100, or 500 mg PBO/kg body weight/day. Three control groups were included. Because a 4-week range-finding study did not provide clear evidence of a NOAEL with respect to minor alterations in liver cell morphology, additional animals were used during the early stages of this study. Thus, after completion of 4 weeks of treatment, 10 males and 10 females from each low-dose group and a special control group consisting of 10 males and 10 females were sacrificed for both gross pathological and histopathologic examinations of the liver. Since the results of the histopathologic examination showed no abnormal findings in the livers of these rats, the low-dose level of 30 mg/kg body weight/day was continued on the 2-year study. The 15-mg/kg body weight/day group was discontinued. No adverse treatment-related effects on survival and no treatment-related clinical signs were seen. A reduced growth rate and a minimal reduction in food intake were noted for males and females receiving 500 mg/kg body weight/day. These females were shown to have increased serum cholesterol levels and slightly higher total serum protein levels. The blood urea nitrogen levels were also slightly higher in this group on one occasion. Dose-related increases in liver and kidney weights were noted in both sexes at 100 and 500 mg/kg body weight/day. Histologically, the most common effect in the liver was centrilobular hepatocyte hypertrophy and the presence of eosinophilic and mixed (basophilic and eosinophilic) cell foci. The severity of focal hyperplasia in the liver was also greater in the intermediate- and high-dose groups. The hyperplastic foci contained either basophilic, normal, or enlarged eosinophilic cells that were variable in size. Normal lobular architecture was retained. Portal triads and central veins were present in the lesions. Neither the incidence of adenomas nor incidence of carcinomas was increased. Pituitary adenomas were common in both males and females but showed no treatment-related effect upon incidence of the adenomas. Thyroid changes, including increased pigment in colloid, and follicular hyperplasia (particularly at 500 mg/kg body weight/day) were seen in both sexes at the end of the study as well as in the high-dose group males dying or sacrificed during the study. In the kidney, glomerulonephritis was more common in the male than the female. No significant increase in the incidence of glomerulonephritis was observed, but the severity of the lesions was increased slightly in the intermediate- and high-dose groups. The non-neoplastic changes observed in this rat study are consistent with induction of the hepatic mixed function oxidase system. The liver observations such as increased liver weight, centrilobular hypertrophy, eosinophilic foci, and eosinophilic focal hyperplasia are probably due to

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enzyme induction. Likewise, the thyroid follicular hyperplasia is a likely secondary response of the thyroid gland to prolonged TSH stimulation resulting from decreased circulating levels of T3 and T4. Reduced T3 and T4 levels are due to increased conjugation and excretion of the thyroid hormones resulting from liver enzyme induction. This pattern has been observed with other liver enzyme inducers such as phenobarbital (McClain, 1989). Following extensive consideration, rat thyroid tumors formed by this mode of action are not considered to be relevant to humans (Cohen et al., 2004; Hill et al., 1999). Given these considerations, the authors concluded that there was no evidence of carcinogenic activity and that the NOAEL based on liver changes was 30 mg/kg body weight/day.

99.3.5.4  Other Species Sarles and Vandergrift (1952) also reported a chronic oral toxicity experiment where a mature female goat was fed a daily dose of 2.0 ml PBO by capsule, 6 days a week, for 1 year. This dose equated to approximately 1000 ppm PBO in the diet. The dose started 4 days after the goat gave birth to a female kid, with both dam and offspring being observed for 1 year to ascertain any signs of direct or indirect (i.e., PBO in dam’s milk) toxic effects. The general health of both dam and kid was unaffected by treatment. The kid was nursed by the treated dam for approximately 6 months and continued to grow and thrive as expected. Red blood cell (RBC) and white blood cell (WBC) counts were unremarkable. At postmortem examination, the dam’s liver revealed slight dystrophy and dysplasia, with central hydrophilic swelling and slight fatty accumulation. No abnormalities were detected in the organs of the kid goat.

99.3.6  Genotoxicity PBO has shown no evidence of mutagenic activity in a number of bacterial assays involving Salmonella typhimurium, Bacillus subtilis, and Escherichia coli both in the presence or absence of rat liver microsomes (S-9) (Ashwood-Smith et al., 1972; Butler et al., 1996; Ishidate et al., 1984; Kawachi et al., 1980; Moriye et al., 1983; White et al., 1977). Most of the studies in systems using mammalian cells in culture show no evidence of mutation or a chromosomedamaging effect. Galloway et al. (1987) investigated a wide range of compounds including PBO in Chinese hamster ovary cells and failed to find chromosome aberrations and sister chromatid exchange in the presence or absence of rat liver S-9. PBO had no effect on Chinese hamster ovary cells in the report of Butler et al. (1996), and produced a small increase in sister chromatid exchanges only in the absence of S-9 in the recent study by Tayama (1996). Tayama also concluded that the metabolites of PBO are unlikely to be genotoxic. In addition, PBO did not produce chromosomal

Hayes’ Handbook of Pesticide Toxicology

aberrations in Chinese hamster lung cells (Ishidate et al., 1984, 1988; Kawachi et al., 1980) or induce mutations in the CHO/HGPT assay (Butler et al., 1996). However, in L5178Y mouse lymphoma cells, PBO showed evidence of mutagenic activity only in the absence of additional metabolic activation. In this study, the mutagenic activity was observed only where cytotox­icity was evident. The relative total cell growth at the lowest mutagenic concentration of 30 g/ml was around 60% (McGregor et al., 1988). PBO also induced cell transformation in Syrian hamster embryo cells (Amacher and Zelljadt, 1983). In this study, three dose levels (0.5, 1.0, and 3.0 g/ml), were used and only two transformed colonies out of 2761 were observed. No indication was given of the dose level that caused the two transformed colonies. Suzuki and Suzuki (1995) investigated the mutagenicity of PBO in human RSA cells, a cell line of double-transformed human embryonal fibroblasts considered to be hyper­mutable, by determining ouabain resistance. The results show an unusual dose response in that despite having little effect upon survival at dose levels above 0.2 g/ml PBO, the inci­dence of mutation declined. The authors also reported mutation of K-ras codon 12 with an apparent similar dose response. While K-ras mutation has been reported in human tumors at various sites (Almoguera et al., 1988; Bos et al., 1987) no such association has been observed in tumors of the rodent liver, the apparent target site of PBO (Maronpot et al., 1995). Butler et al. (1996) reported that PBO did not induce unscheduled DNA synthesis in rat hepatocytes. Moreover, Beamand et al. (1996) showed a similar negative response to PBO in cultured human hepatocytes. In vivo studies have also failed to demonstrate convin­ cing genotoxic effects of PBO. A dominant lethal assay in ICR/Ha Swiss mice using both single and multiple doses of PBO given either by intraperitoneal injection or by gavage resulted in toxicity and death of the male mice. Although there was some evidence of reduced reproductive efficiency and an increase in early fetal death, the results were not consistent and the authors concluded the study was equivocal (Epstein et al., 1972). Other studies have been reported only briefly as abstracts and have stated that no chromosomal aberrations or sister chromatid exchanges were produced in either rat or mouse bone marrow (Ivett and Tice, 1983; Kawachi et al., 1980).

99.3.7  Mode of Action Considerations   for Oncogenicity The chronic toxicity/oncogenicity studies discussed in Section 99.3.4 indicate that liver is a target for both oncogenic and nononcogenic changes in both the mouse and the rat. In addition, one study showed an apparent hyperplastic response in the thyroid (Butler et al., 1998).

Chapter | 99  The Safety Assessment of Piperonyl Butoxide

Many nongenotoxic compounds have been shown to induce tumors in rodents and consideration of their relevance to humans has been the focus of intense effort over the past two decades (Ames et al., 1993; Butler, 1996; Cohen and Ellwein, 1990; Cohen, et al., 2003, 2004; Grasso and Hinton, 1991; Grasso et al., 1991; Holsapple et al., 2006; Loury et al., 1987; Wilson et al., 1992). Much progress has been made in understanding the mode of action by which liver by nongenotoxic inducers of hepatic xenobiotic metabolism, such as sodium phenobarbital (NaPB) and PBO, may produce liver tumors (Grasso and Hinton, 1991; Grasso et al., 1991; Holsapple et al., 2006). As described above, it appears that two distinct responses are occurring during the lifetime dietary administration of PBO. Clear evidence of carcinogenicity is observed at very high, toxic doses leading to liver tumors in rats and mice (Takahashi et al., 1994a,b, 1997). Studies show that benign liver tumors may be produced in the mouse (but not the rat) at the MTD, likely by a mode of action (MOA) different from what is occurring at the excessively high doses used in the studies by Takahashi et al. (1994a,b, 1997). It is important to note that other studies show the lack of hepatocarcinogenicity of PBO in the mouse and rat at MTD and lower doses. The PBO-induced mouse liver tumors are likely due to a MOA similar to that described for phenobarbital and related compounds (Holsapple et al., 2006). Phenobarbital is known to induce liver tumors in rodents (Whysner et al., 1996). A diagnostic effect of phenobarbital in rodent liver is the induction of CYP forms, particularly of CYP2B forms (Okey, 1990; Nims and Lubet, 1996; Parkinson, 2001). The pleiotropic effects of phenobarbital in rodent liver including increased liver weight, liver hypertrophy, increased replicative DNA synthesis, induction of CYP forms, and liver tumor promotion are mediated through various nuclear receptors, particularly the CAR (Dickins, 2004; Honkakoski et al., 2000; Tien and Negishi, 2006; Ueda et al., 2000). Indeed, in mice lacking CAR, phenobarbital does not induce CYP2B forms, does not increase liver weight, and does not stimulate replicative DNA synthesis (Wei et al., 2000; Yamamoto et al., 2004). Moreover, while phenobarbital promoted liver tumors in normal mice given a single dose of diethylnitrosamine, no hepatocellular adenomas or carcinomas were observed in mice lacking CAR (Yamamoto et al., 2004). Several studies have shown that PBO is an inducer of hepatic xenobiotic metabolism in the mouse and rat (Fennell et al., 1980; Goldstein et al., 1973; Lake et al., 1973; Phillips et al., 1997; Wagstaff and Short, 1971). Phillips et al. (1997) characterized the induction of enzyme activ­ ities in both the mouse and rat and compared it to the classic inducer, phenobarbital. Four groups of 16 male F-344 rats were fed PBO in the diet at 100, 550, 1050, or 1850 mg/kg body weight/day. Animals were treated for either 7 or 42 days and were sacrificed, and liver studies

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were performed. Even the low-dose group (100 mg/kg body weight/day PBO) showed increased relative liver weights and microsomal protein content (42 days treatment), increased cytochrome P450 levels (7 days treatment), increased GGT levels (42 days treatment), and increases in certain MFO enzyme activities. PBO appeared to be a mixed-type enzyme inducer in the rat in that it induced hepatic cytochrome P450 isoenzymes in the CYP1A, CYP2B, and CYP3A subfamilies. Recent work by Watanabe et al. (1998) in the rat is in agreement with these findings of Phillips. These workers also noted weak induction of the CYP4A isozyme. The induction pattern was similar to that they observed for NaPB, with the exception that NaPB did not induce CYP1A. A NOEL of 0.05% PBO in the diet (approximately 50 mg/kg body weight) over 4 weeks was reported for this enzyme induction. PBO treatment of the mouse by Phillips and co-workers also resulted in a dose-related induction of cytochrome P450 content and ethylmorphine demethylase activity (CYP3A). Taken with other studies that have shown induction of CYP1A and CYP2B isoenzymes in mouse liver (Adams et al., 1993; Fennell et al., 1980), PBO appears to be able to induce CYP1A, CYP2B, and CYP3A iso­ enzymes in CD-1 mouse liver as well. Aside from the enzyme induction discussed above, certain nongenotoxic rodent liver carcinogens produce either a transient or a sustained stimulation of cell replication (Goldsworthy et al., 1991). Both PBO and NaPB produced a stimulation of cell replication after 7 but not 42 days of treatment in the mouse (Phillips et al., 1997). Like NaPB, PBO also increased relative liver weight in CD-1 mice and produced liver hypertrophy, although a difference in the lobular distribution of this effect was noted. Generally, the effects of PBO on relative liver weight, liver morphology, replicative DNA synthesis, and xenobiotic metabolism occurred at the 100 and 300 mg/kg body weight/day dose levels, the same doses where eosinophilic nodules were observed in a 2-year study by Butler et al. (1998). In addition, the eosinophilic nodules produced by PBO in mouse liver (Butler, 1996) appear similar to those formed by NaPB (Evans et al., 1992). Many studies have demonstrated that PBO can induce CYP forms and CYP-dependent enzyme activities in mice and rats (Adams et al., 1993a,b; Goldstein et al., 1973; Hodgson and Philpot, 1974; Hodgson et al., 1995; Fennell et al., 1980; Lake et al., 1973; Mugumura et al., 2006; Phillips et al., 1997; Ryu et al., 1996, 1997; Wagstaff and Short, 1971; Watanabe et al., 1998). The induction of CYP forms has been evaluated by the measurement of enzyme activities, mRNA levels, and CYP apoprotein levels. Studies in the mouse have demonstrated that PBO can induce hepatic CYP2B forms and also other CYP subfamily forms including CYP1A forms. These studies have been performed in various mouse strains including strains both responsive (e.g., C57BL/6) and nonresponsive

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(e.g., DBA/2) to CYP1A induction by aromatic hydrocarbons. Male C57BL/6 N mice were given a single intraperitoneal injection of 338 mg/kg (1 mmol/kg) PBO and killed 24 h later (Adams et al., 1993a). Treatment with PBO significantly induced hepatic microsomal 7-ethoxyresorufin Odeethylase, acetanilide hydroxylase and 7-pentoxyresorufin O-depentylase activities, which were employed as enzymatic markers of CYP1A1, CYP1A2, and CYP2B10, respectively. Western immunoblotting of liver microsomes revealed an induction of CYP1A1, CYP1A2, and CYP2B10 apoproteins, whereas while PBO induced CYP1A2 and CYP2B10 mRNA levels, only a weak induction of CYP1A1 mRNA was observed. In another study, male C57BL/6 mice were given single intraperitoneal doses of 52, 104, 156, 208, and 416 mg/kg PBO and killed 24 h later (Adams et al., 1993b). Treatment with PBO induced hepatic microsomal 7-ethoxyresorufin O-deethylase, acetanilide hydroxylase and 7-pentoxyresorufin O-depentylase activities, CYP1A1, CYP1A2, and CYP2B10 mRNA levels, and CYP1A1, CYP1A2, and CYP2B10 apoprotein levels by Western immunoblotting. The effect of PBO on enzyme activities and mRNA and apoprotein levels was generally dose-dependent. Male C57BL/6 and DBA/2 mice were given single intraperitoneal 254 and 508 mg/kg (0.75 and 1.5 mmol/kg) doses of PBO and killed 24 h later. While CYP1A2 and CYP2B10 apoprotein levels were induced by PBO in both mouse strains, CYP1A1 apoprotein levels were only induced in liver microsomes from C57BL/6 mice (Adams et al., 1993b). Male C57BL/6 and DBA/2 mice were given a single intraperitoneal injection of 400 mg/kg PBO and killed 24 h later (Ryu et al., 1997). PBO induced CYP1A1 mRNA in both mouse strains, the effect being more marked in C57BL/6 mice. Treatment with PBO also induced CYP1A2 mRNA levels and immunocytochemical studies demonstrated a centrilobular induction of CYP1A1/2 in liver sections from both mouse strains. Additional studies in aryl hydrocarbon receptor (AhR) knockout mice have confirmed that AhR-independent pathway(s) are involved in induction of CYP1A2 by PBO (Ryu et al., 1996). The treatment of male ICR mice with 6000 ppm PBO in the diet for 1, 4, and 8 weeks significantly elevated CYP1A1, CYP2A5, CYP2B9, and CYP2B10 mRNA levels (Muguruma et al., 2006). The treatment of male C57BL/10 and DBA/2 mice with 500 mg/kg/day PBO by intraperitoneal injection induced some hepatic microsomal CYP-dependent enzyme activities (Fennell et al., 1980), whereas in another study the treatment of male CD-1 mice with 100 and 300 mg/kg/day PBO by dietary administration for 42 days resulted in increased microsomal total CYP content (Phillips et al., 1997). Male F344 rats were given 0 (control), 100, 550, 1050, and 1850 mg/kg/day PBO by dietary administration for periods of 7 and 42 days (Phillips et al., 1997). With the exception of the lowest PBO dose level (where

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the changes were not always statistically significant), significant increases in microsomal total CYP content and 7-ethoxyresorufin O-deethylase, 7-pentoxyresorufin Odepentylase and ethylmorphine N-demethylase activities were observed at both time points. The three enzyme activities are markers for induction of CYP1A, CYP2B, and CYP3A/other CYP forms, respectively. In another study, male F344 rats were treated with diets containing 0.05, 0.2, and 2% PBO for 1 and 4 weeks (Watanabe et al., 1998). Significant increases in microsomal total CYP content, some CYP-dependent enzyme activities, glutathione S-transferase activities, and one UDP glucuronosyltransferase activity were observed at the highest PBO dietary level after 1 and 4 weeks, with some increases being observed in rats given 0.2% PBO, whereas no significant effects were observed in rats given 0.05% PBO. Western immunoblotting studies were conducted with liver microsomes from control rats and rats given 2% PBO for 4 weeks. Treatment with PBO resulted in an induction of CYP1A1, CYP2B1/2, and CYP3A apoprotein lev­els, with a weak induction of CYP4A apoprotein levels also being reported (Watanabe et al., 1998). Finally, immunocytochemical procedures were employed to demonstrate a centrilobular induction of CYP1A1/2 and CYP2B1/2 in liver sections from rats given 2% PBO for 4 weeks. The treatment of male Wistar rats with 1200 mg/kg/day PBO by oral administration for up to 7 days was reported to upregulate CYP genes in the 1, 2, 3, and 4 families (EllingerZiegelbauer et al., 2005). With respect to the rat, Watanabe et al. (1998) demonstrated the production of centrilobular hypertrophy in the rat liver following 4 weeks of dosing with 2% PBO in the diet. The degree of response was similar to that observed after 4 weeks dosing with 0.1% NaPB. Further evidence of the similarity between the action of PBO and NaPB comes from a report by Okamiya et al. (1998) showing an increase in proliferating cell nuclear antigen in rats fed 0.2% PBO in the diet for 4 weeks. They also demonstrated a decrease in the gap junction protein connexin 32 (Cx32) at the highest dose tested (2.0%) after dosing for 1, 2, but not 4 weeks. Phillips et al. (1997) reported that high doses of PBO (1850 mg/kg body weight/day) given to rats caused a significant reduction of body weight gain and of food consumption throughout the 42 days of dosing. Morphological examination of liver showed individual cell necrosis in rats dosed with 1050 and 1850 mg PBO/kg body weight/ day. While the severity of the individual cell necrosis was similar in rats given 1050 and 1850 mg PBO/kg body weight/day, the incidence was greater (seven of eight animals examined) in rats given 1850 mg/kg body weight/day. Like NaPB, the increase in relative liver weight in PBOtreated animals was associated with hypertrophy, although a difference in the lobular distribution of this effect was noted. Replicative DNA synthesis was stimulated by 550

Chapter | 99  The Safety Assessment of Piperonyl Butoxide

and 1050 mg PBO/kg body weight/day and 0.05% NaPB after 7 days of administration, most likely due to transient mitogenesis typical of enzyme inducers. In contrast, the stimulation of cell replication observed after 42 days treatment by 1050 mg/kg body weight/day is more likely to be associated with the onset of a regenerative hyperplasia. The mitogenic and hypertrophic effects of PBO were observed at doses (e.g., 550 mg/kg body weight/day) lower than those required to produce individual cell necrosis, whereas a high incidence of necrosis was only observed in rats given 1850 mg/kg body weight/day for 42 days. Chronic treatment (i.e., 42 days) with PBO at high-dose levels such as that in an oncogenesis study could result in a sustained stimulation of replicative DNA synthesis and an increased likelihood of oncogenesis. It is important to note that Takahashi et al. (1994b) reported no increase in liver tumors following 2 years of dosing at the lowest study dose of 547 mg/kg body weight/day, a dose calculated by the authors to be about 18,000 times the allowable daily intake (ADI) for humans. In contrast, higher doses of PBO were both toxic to the rat liver and produced tumors. Both of these mechanisms are threshold phenomena and suggest that at doses likely to be encountered by humans, PBO poses essentially no oncogenic risk.

99.3.8  Human Studies Wintersteiger and Juan (1991) investigated the absorption of combination pyrethrins and PBO sprays across the skin of six healthy subjects in Austria. The spray was applied over a wide area of the back with a total dose of approximately 3.3 mg pyrethrum extract and 13.2 mg PBO being administered. No untoward clinical signs were noted. Cutaneous absorption of PBO was shown to be extremely low, with plasma samples containing no more than 10 ng PBO/ml. Wester et al. (1994) investigated the percutaneous absorption of both PBO and pyrethrin compounds across the skin of the ventral forearm in six volunteers. Based on the recovery of radioactivity in the urine, it was calculated that 2.1%  0.6% of the dose of PBO was absorbed through the skin. Not surprisingly, higher levels of absorption of PBO were achieved when the compound was applied to the skin of the scalp (8.3%). There was no evidence of any local or systemic toxicity of PBO when used as a topical agent in humans. The most definitive study on the absorption and excretion of PBO was a mass balance study reported by Selim (1995) using 14C PBO from two different formulations following dermal application to healthy volunteers. The first preparation applied was a 4% (w/w) solution of PBO in an aqueous formulation. This product was applied to four healthy human volunteers. The second preparation tested was a 3% (w/w) solution of PBO in isopropyl alcohol. In the former case, the mean amount of PBO applied was 3.8 mg per volunteer (approximately 39.9 Ci of radioactivity per volunteer), while the average exposure was 3.0 mg

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PBO per volunteer (approximately 40 Ci of radioactivity per volunteer) in the case of the isopropyl alcohol solution. Results from this study show that there was a similar dermal absorption pattern for both formulations. The principal route of excretion of absorbed radioactivity was via urine. Fecal samples from volunteers contained negligible levels of radioactivity. Dermally applied PBO was rapidly excreted from the volunteers. The majority of the applied radioactivity remained at the application site, with less than 3% of the applied dose being absorbed during the 8-h test period. Radioactivity did not accumulate in the skin. The only study to examine the ability of PBO to inhibit xenobiotic metabolism in humans was reported by Conney et al. (1972). Using the rate of antipyrine metabolism as a gauge of P450 activity, two healthy men (weighing 87.4 kg and 82.6 kg, respectively) were given capsules containing increasing amounts (5, 10, 20, and 50 mg) of PBO at consecutive intervals of approximately 1 week. No signs of toxicity were seen. Clinical chemical analysis of blood and urine samples taken at 4, 8, and 24 h after ingestion of the capsules did not reveal any adverse effects. In a subsequent study, eight men were given PBO as a single dose of 0.71 mg/kg body weight. A control group received a placebo. Two hours later, both the treated and control groups received a 250-mg oral dose of antipyrine. Antipyrine was analyzed in blood samples taken at intervals over the next 31 h. PBO had no effect on the rate of clearance of antipyrine. Although no systematic epidemiology studies have been conducted on PBO-exposed individuals, no evidence sug­gests that PBO has resulted in any significant adverse effects to human health. Occupational health data from a PBO manufacturing site in Italy, consisting of routine health checks, were collected on potentially exposed workers from 1974 to 1994. Sixty workers were examined, 11 of whom were employed in the manufacture of PBO for periods of 15–26 years. Workers received x-rays and spirometric evaluations every 3 years and annual clinical chemistry analysis. No adverse clinical signs or symptoms related to PBO were found (Endura, 1996). Similarly, at a manufacturing site in Scotland, where PBO was made from 1962 to 1990, “no cases of toxic symptoms or adverse effects attributable to PBO manufacture” were noted in production workers at the plant. Moreover, no adverse effects were reported in operations involved in the handling and use of PBO at a site in England (JMPR, 1993; Pitman Moore, 1990; Wellcome, 1991). The only clinical report referring to PBO exposure is the case of two sisters who gave birth, within 2 weeks of each other, to children who each had coarction of the aorta (Hall et al., 1975). Both mothers had been on a camping trip at 2 months gestation where they used “large amounts” of insect repellents and insecticides containing, among other chemicals, PBO, pyrethrins, DEET, and the organophosphorus insecticide DDVP. No cause-and-effect relationship was established.

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There have been no reported suicide attempts by humans with PBO. Based on its acute toxicity to experimental animals, a probable oral lethal dose for humans is estimated to be 5–15 g/kg body weight (i.e., approximately 300–900 g PBO for a 60-kg human).

The authors conclude that in view of their widespread use, the data indicate that PY/PBO products can be used with a relatively low risk of adverse effects. Moreover, the data suggest that they are not likely to cause reactions in people with asthma or allergies.

99.3.9  Human Experience

99.4  Pharmacodynamics

As mentioned earlier, although PBO is not used alone, it has been used extensively in combination with the pyrethrins. Insight into the human experience with pyrethrins and PBO-containing products comes from an investigation of human incidents reported through the American Association of Poison Control Centers (AAPCC) Toxic Exposure Surveillance System (TESS) associated with regulated insecticides containing pyrethrins and PBO (PY/PBO) from 2001 to 2003 (Osimitz et al., 2009). Sales of household insecticides containing PY/PBO during the period of this review (2001–2003) were estimated to be over 40,000,000 individual product units. Assuming that a unit is used on four occasions, this suggests that over 160,000,000 uses of PY/PBO products occurred during the review period. Given that many households have more than one person in them, it is reasonable to assume that perhaps over 300,000,000 people may have had some exposure to PY/PBO products over the 3 years of the study period. Although tabulation of incidents included all pyrethrins/ PBO exposures regardless of effect, the focus was placed on those incidents reporting potential dermal and respiratory effects. Because the TESS system does not contain a data field designated to capture the caller’s allergy or preexisting medical condition information, such as allergy status of hypersensitivities, a surrogate measure of association was selected. The treatment options documented in each case record were used to identify treatments typically associated with more intense respiratory or dermal effects. Records were searched for indication of bronchodilator use as part of the treatment, suggesting that simple fresh air or ventilation was insufficient to alleviate signs and symptoms likely associated with shortness of breath in patients with preexisting airway disease. With respect to dermal effects, steroid treatment as suggestive of a need for treatment involving more than simple irrigation or washing of the affected area was reviewed. The limitations notwithstanding, the analysis showed that: Despite extensive use, incidents with reports of moderate or major adverse effects were relatively rare (717 moderate and 23 major outcomes out of 17,873 calls). l Following label-directed use of the products, adverse dermal or respiratory reactions were very rare; (dermal: 17 moderate, 1 major; respiratory: 18 moderate, 0 major). l The data suggest that asthmatics and people sensitive to ragweed (Ambrosia artisifolia) are not unusually sensitive to PY/PBO. l

Yamamoto (1973) suggests that the primary function of synergists such as PBO when formulated with pyrethrins or pyrethroids is to provide an alternative substrate for the MFO enzyme system, which would normally metabolize such insecticides. Inhibition of MFO-mediated oxidation of the transmethyl groups and the alcohol moiety on the pyrethrin molecule appear to be the most important functions of PBO. Inhibition of ester hydrolysis may also contribute to the effectiveness of PBO as a synergist. As a known alternative substrate for the liver microsomal enzyme system, PBO will inhibit the metabolism of many xenobiotics including drugs and pesticides. Brown (1970) reported that the detoxification of certain drugs such as pentobarbital, zoxazolamine, antipyrine, and benzopyrene were inhibited by PBO, presumably due to the inhibition of their microsomal oxidation. Conney et al. (1972) investigated the inhibition of antipyrine metabolism in rats and mice. Both species were treated intraperitoneally (i.p.) with a single dose of PBO, followed by a further i.p. injection of antipyrine (200 mg/kg body weight) 1 h later. A marked species difference was noted in the response; the no-observable-effect level (NOEL) for inhib­ ition of antipyrine metabolism in the mouse was 0.5–1.0 mg PBO/kg body weight, whereas the NOEL for the rat was 100 mg PBO/kg body weight. The effects of PBO on the metabolism of benzopyrene in Sprague-Dawley rats were studied by Falk et al. (1965). PBO was administered by the oral, intraperitoneal (i.p.), or intravenous (i.v.) routes at various times before the i.v. injection of labeled benzopyrene. The level of radioactivity was then measured in bile at frequent intervals up to 4 h. This author demonstrated marked inhibition of benzopyrene metabolism when PBO was administered i.v. at 262 mg/kg body weight, some 5 min to 16 h before the administration of benzopyrene. However, this effect is much reduced at 121 mg/kg body weight. Virtually no effect was seen at 25 h post-dosing. This implies that single large doses of PBO are quickly metabolized by rats. Administration of PBO by the oral and i.p. routes resulted in a greatly reduced effect when compared with the i.v. route. A second similar study carried out by Conney et al. (1972), where the effects of i.p. administration of PBO on the metabolism of benzopyrene were investigated, showed less sensitive results when rats of lighter weight were used (approximately 180 g versus 400 g in Falk’s study). It was postulated by Brown (1970) that the extra fat in the

Chapter | 99  The Safety Assessment of Piperonyl Butoxide

animals in Falk’s study could possibly act as a reservoir of PBO and lead to a longer duration of action. It would appear from these studies in rats that 250 mg of PBO per kg body weight is the minimum oral dose required to give any significant effect on benzopyrene metabolism. Whereas a single dose of PBO will generally inhibit the metabolism of pentobarbital, repeated PBO doses will generally induce the metabolism of phenobarbital and other xenobiotics. Brown (1970) reports an experiment in rats where a single i.p. dose of PBO (333–1000 mg/kg body weight) increased the sleeping time of the animals following administration of pentobarbital. However, the i.p. administration of eight injections of 50 mg PBO/kg body weight, each at 12-h intervals, followed by the injection of pentobarbital some 18 h later, caused a reduction in sleeping time in rats. The administration of 50 mg PBO per kg body weight i.p. to rats (Anders, 1968) and mice (Graham et al., 1970) prior to treatment with hexobarbital approximately doubled the sleeping time of both species. CD-1 mice given a single i.p. dose of 600 mg PBO per kg body weight were found to have suffered less hepatotoxicity when treated with acetaminophen (600 mg/kg body weight, p.o.) at either 2 h prior to or 1 h following PBO administration. This reduced hepatotoxicity was measured via GSH and sorbitol dehydrogenase levels, as well as subsequent histo­ pathology of the liver. Since the hepatic MFO system metabolizes acetaminophen to a toxic metabolite, the decreased toxicity seen in this experiment is likely due to inhibition of such oxidase enzymes by PBO (Brady et al., 1988). Many other studies have been undertaken relevant to the pharmacodynamics of PBO and are reported elsewhere (Skrinijaric-Spoljar et al., 1971; Conney et al., 1972; Goldstein et al., 1973). More recent work is discussed in Section 99.3.5 of this chapter.

99.5  Exposure assessment 99.5.1  Dietary Exposure The first formal discussion of human exposure to PBO took place at a joint FAO/WHO Codex Alimentarius in 1987. FAO/WHO estimated that the average daily human diet (1.4 kg) might contain as much as 1 ppm of PBO, corresponding to a daily dose of 1.4 mg. This is likely an over­estimation as cooking may destroy up to 90% of PBO present. In a refinement of this exposure estimate, Crampton (1994) calculated that the daily exposure of adults to PBO residues in food was 0.0037 mg/kg body weight/day. During the 1990s, the PBO Task Force (PBTF), a consortium of PBO producers and marketers, developed extensive exposure data for PBO for use in dietary exposure assessments. In addition to a series of crop residue studies on various crops representing 12 crop groups, studies were conducted where residue levels and transfer factors were obtained following application of PBO to livestock.

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The USPEA used these data, when appropriate, along with two models, the Lifeline model (Version 2.0) and the Dietary Exposure Evaluation Model software with the Food Commodity Intake Database (DEEM-FCID, Version 2.03) to estimate acute and chronic dietary exposure to PBO in the United States. Both models incorporated food consumption data from the USDA’s Continuing Surveys of Food Intakes by Individuals (CSFII) from 1994 to 1996 and 1998. USDA Pesticide Data Program (PDP) data were used for commodities that have pre-harvest registered uses and for cereal grain crops that have a stored grain use. All other commodities were assigned residues from either a simulated warehouse space spray experiment or a simulated restaurant experiment. Residue data from studies conducted following dermal treatment of livestock were used as input values for meat, milk, poultry, and eggs. The residues of concern for plants include the parent PBO and a twofold factor to account for metabolites, unless field trial data for metabolites on related crops indicated a lower factor was appropriate. Percent crop treated data were used for all commodities for which percent crop treated data are available. Where no percent crop treated data were available, the dietary analyses assumed 100% crop-treated. Based on the methodology discussed above, the acute dietary exposure at the 99.9th percentile was estimated as 0.2865 mg/kg/day and 0.3467 mg/kg/day for adults (50  years) and females (13–49 years), respectively. Chronic dietary exposures were estimated at 0.0066 mg/kg/day and 0.0070 mg/kg/day for adults (50  years) and females (13–49 years), respectively. The dietary exposure estimates likely significantly overestimate real exposure. Important refinements include obtaining additional residue data and additional percent crop-treated information

99.5.2  Non-Dietary Exposure FAO/WHO estimated the exposure from aerosol consumer products (from ingestion and inhalation) to be approximately 0.63 mg/day (FAO/WHO, 1987). Given the extensive non-dietary use of PBO, the NonDietary Exposure Task Force (NDETF) was established in 1996 to develop a long-term program to conduct a series of transferability studies to better understand the phenomenon of human exposure to pesticides used in the home. Most of the studies were conducted with formulations of pyrethrins/PBO and permethrin/PBO and focused on the use of fogger and aerosol products indoors. Carpet and vinyl were selected as the flooring surfaces of interest because of their different physical and chemical properties and because they represent a significant amount of the floor coverings used in homes in North America. While the focus of the NDETF efforts was on total release foggers, a study was also conducted to determine both dispersion (air levels) and deposition (on flooring) of pyrethrin/PBO resulting from the use of a hand held aerosol spray can.

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Potential direct exposure of the user was also measured. Air sampling from the breathing zone of the applicator and analysis of residues on cotton gloves was performed. The Food Quality Protection Act (FQPA) requires that the U.S. EPA consider available information concerning the cumulative effects of a particular pesticide and other substances that have a common mechanism of toxicity. The U.S. EPA has confirmed that PBO does not appear to produce a toxic metabolite produced by other chemicals nor does it share a common mechanism of action with others. Using the data developed by the NDETF as well as other sources, they estimated exposures for the following scenarios (U.S. EPA, 2005a): Handler Liquid spray formulation by low-pressure handwand for indoor surface spray application Liquid spray formulation by low-pressure handwand for indoor crack and crevice treatment Liquid spray formulation by hose-end sprayer for lawn and garden application l Postapplication Inhalation exposure from airborne application of mosquito adulticide Inhalation exposure from application of mosquito adulticide from truck mounted sprayer Toddler incidental ingestion of residue from treated turf grass via hand-to-mouth activities Toddler incidental ingestion of residue via object-tomouth activity while on treated grass Toddler incidental ingestion of soil from treated area Toddler incidental ingestion of residues deposited on carpet via hand-to-mouth activities after use of total release foggers Toddler incidental ingestion of residues deposited on vinyl flooring via hand-to-mouth activities after use of total release foggers Toddler incidental ingestion of residues on pets via hand-to-mouth activities after pet treatment Inhalation exposure to aerosol spray during and after space spray application The details of these assessments are beyond the scope of this chapter, but estimated exposures ranged from 0.000079 mg/kg/day (handler exposure while using hoseend sprayer, short- and intermediate-term exposure) to 0.28 mg/kg/day (short-term child ingestion of residues on pets via hand-to-mouth activities after pet treatment). l

99.6  Risk characterization 99.6.1  Cancer The Joint FAO/WHO Meeting on Pesticide Residues (JMPR) evaluated the toxicology of PBO in 1965, 1966, 1972, 1992,

Hayes’ Handbook of Pesticide Toxicology

and, most recently in 1995. They concluded that, at doses up to internationally accepted standards for a Maximum Tolerated Dose, PBO is not oncogenic in the mouse or rat. Based on this assessment, no risk assessment is warranted (JMPR, 1995). The U.S. EPA evaluated the weight of evidence relating to the potential oncogenicity of PBO and classified it as a Group C-Possible Human Carcinogen (U.S. EPA, 1995a). This was based on the increases in hepatocellular tumors in both male and female mice (adenomas, carcinomas, and combined adenomas and carcinomas in the males and adenomas only in the females). However, because of the generally low concern for mutagenicity, and the minor significance of other tumors observed in the rat studies, rather than recommend the Q*1 linearized multistage model for risk characterization based on oncogenicity, the U.S. EPA endorsed the use of a Reference Dose (RfD) and Margin of Exposure (MOE) approaches using nononcogenic endpoints, such as body weight changes. Although, as discussed previously, the data suggest that the MOA resulting in PBO-induced liver tumors in mice is not plausible in humans, quantitative differences exist between the doses of PBO that produce liver tumors in mice and human exposure to PBO. In the PBO mouse bioassay showing liver tumors male mice were given 100 and 300 mg/kg/day PBO and female mice were given 300 mg/ kg/day PBO (Butler et al., 1996). The lowest dose level at which tumors were seen was thus 100 mg/kg/day. The proposed Acceptable Daily Intake (ADI) of PBO established by the Joint FAO/WHO Meeting on Pesticide Residues in 1995 is 0.2 mg/kg/day (JMPR, 1995). The U.S. EPA has estimated the Population Adjusted Dose (PAD) for acute human PBO exposure to be 6.3 mg/kg/day and chronic human exposure to be 0.155 mg/kg/day (U.S. EPA, 2005b) (The PAD is the Acute reference dose (RfD) or the Chronic RfD modified by the FQPA Safety Factor. The safety factor for both the acute and chronic dietary assessments of PBO is 1X.) The lowest dose level for liver tumors in mice (100 mg/kg/day) is 645 times greater than the “permitted” level for human chronic exposures to PBO. Actual expos­ ure is much less: the U.S. EPA estimates chronic dietary exposure to PBO for the general population to be approximately 0.008 mg/kg/day. Thus, the lowest dose level that was associated with liver tumors in mice (100 mg/kg/day) is approximately 12,500 times greater than the estimated level for human chronic exposures to PBO (U.S. EPA, 2005b).

99.6.2  Non-Cancer Effects As mentioned above, the Joint FAO/WHO Meeting on Pesticide Residues (JMPR) evaluated the toxicology of PBO in 1965, 1966, 1972, 1992, and, most recently in 1995. Based on the NOAEL of 600 ppm (16 mg/kg body weight/day) in the most sensitive toxicology study (1 year feeding study in dogs, Goldenthal, 1993b), they established

Chapter | 99  The Safety Assessment of Piperonyl Butoxide

an Allowable Daily Intake (ADI) for humans of 0.2 mg PBO/kg body weight (JMPR, 1995). Following a careful evaluation of the hazard data, the U.S. EPA, as part of the reregistration of PBO, identified a series of toxicological endpoints used to estimate risk using the Margin of Exposure (MOE) approach. They concluded that no quantitative dermal assessment was required because no systemic effects were observed at the limit dose (1000 mg/kg/day) in the 21-day dermal absorption study in rabbits. The EPA selected benchmark studies for acute inhalation exposure, short-, intermediate- and long-term inhalation exposure, and short- and intermediate-term incidental oral exposure. MOE levels of concern ranged from 100 to 1000, meaning that MOEs below these values may trigger risk management measures. The results of the nondietary risk assessments indicate that all residential ex­posure scenarios result in MOEs greater than the applicable MOE levels of concern and thus are of no safety concern.

99.6.2.1  Dietary Risk Dietary risk assessment incorporates both exposure to and toxicity of a given pesticide. Both the FAO/WHO estimate of dietary exposure of 1.4 mg/person/day (0.02 mg/ kg/day) and the Crampton (1994) estimate of daily expos­ ure of adults to PBO residues in food of 0.0037 mg/kg body weight/day are much less than the ADI of 0.2 mg/kg/day. The U.S. EPA expresses dietary risk as a percentage of a level of concern (referred to as the population adjusted dose, or PAD). The PAD is the dose predicted to result in no unreasonable adverse health effects to humans, including sensitive subgroups. Thus, it is a function of the reference dose (RfD) and the appropriate FQPA safety factor. Exposures less than 100% of the PAD are below EPA’s level of concern. For acute dietary exposure, the EPA used the no-observed-adverse-effect-level (NOAEL) of 630 mg/ kg/day from a rat oral developmental toxicology study (Tanaka et al.,1995). The aPAD was calculated as 630 mg/ kg/day ÷ 100 safety factor  6.3 mg/kg/day. Exposure estimates for the U.S. population and the highest exposure group (children 1–2 years old) show exposures at 6% and 20% of the aPAD, respectively. The chronic dietary endpoint came from a dog study with a NOAEL of 15.5 mg/kg/day (Goldenthal, 1993b). The exposure estimate for the U.S. population is 5% of the cPAD and 12% for the highest exposed subpopulation, children (1–2 years of age).

99.6.2.2  Non-dietary Risk The FAO/WHO estimate of 0.63 mg/day exposure from aerosol consumer products corresponding to a daily dose of 0.009 mg/kg body weight for a 70-kg adult and 0.042 mg/ kg body weight for a 15-kg child is well below the ADI

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for humans of 0.2 mg/kg body weight for PBO set by the JMPR (1995). The more recent and more highly refined assessment by the U.S. EPA considered three product user (handler) and nine postapplication residential exposure scenarios. The results of the residential exposure assessment indicated that all residential exposure scenarios assessed showed MOEs greater than the applicable target MOEs (ranging from 600 for a child playing with a pet treated with spray to more than 1,000,000 for incidental ingestion risks to toddlers reentering treated lawns). All residential scenarios result in exposures below the level of concern.

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Chapter | 99  The Safety Assessment of Piperonyl Butoxide

U.S. Environmental Protection Agency (1998). “Assessment of Thyroid Follicular Cell Tumors,” EPA/630/R-97/002. U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (2005a). “Occupational and Resi­dential Exposure Assessment and Recommendations for the Reregis­ tration Eligibility Decision (RED) For Piperonyl Butoxide,” Office Of Prevention, Pesticides And Toxic Substances, Washington, DC. U.S. Environmental Protection Agency (2005b). “Piperonyl Butoxide HED Risk Assessment for Reregistration Eligibility Document (RED),” Office Of Prevention, Pesticides And Toxic Substances, Washington, DC. U.S. National Cancer Institute (1979). “Bioassay of PBO for Possible Carcinogenicity,” DHEW Publ. 79–1375. U.S. Department of Health, Education and Welfare, Bethesda, MD. Van Miller, J. P., Lalich, J. J., and Allen, J. R. (1977). Increased incidence of neoplasms in rats exposed to low levels of 2,3,7,8-tetrachlorodibenzo-rhodioxin. Chemosphere 6(9), 537–544. Varsano, R., Rabinowitch, H. D., and Rubin, B. (1992). Mode of action of piperonyl butoxide as herbicide synergist of atrazine and terbutryn in maize. Pestic. Biochem. Physiol. 44, 174–182. Wachs, H. (1947). Synergistic insecticides. Science 105, 397–401. Wagstaff, D. J., and Short, C. R. (1971). Induction of hepatic microsomal hydroxylating enzymes by technical PBO and some of its analogues. Toxicol. Appl. Pharmacol. 19, 54–61. Watanabe, T., Manabe, S., Ohashi, Y., Okamiya, H., Onodera, H., and Mitsumori, K. (1998). Comparison of the induction profile of hepatic drug-metabolizing enzymes between PBO and phenobarbital in rats. J. Toxicol. Pathol. 11, 1–10.

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Chapter 100

Rodenticides Alain F. Pelfrène Charbonnières-les-Bains, France

100.1  Introduction Rats and mice compete with humans for food. This loss to rodents causes economic loss everywhere. In some developing countries it can cause starvation. Rodents are also hosts for human diseases, including plague, endemic rickettsiosis, leishmaniasis, spirochetosis, tularemia, leptospirosis, tick-borne encephalitis, and listeriosis. Rats occasionally bite people. Finally, rodents do a variety of other damage, mainly by gnawing. Insofar as possible, rodent populations should be controlled by limiting their access to food and harborage. Individual animals or small groups may be removed conveniently by trapping. However, there will always be a need for poisons in rodent control. Unfortunately, effective permanent control through poisoning is not simple. The animals must be enticed to ingest a toxicant in sufficient dosage if the effort is to succeed. But rodents rarely or ever constitute an important problem unless they have a supply of food and water. This means that, in spite of containing a foreign substance, the solid or liquid bait used should be at least as attractive to the rodents as their usual supply of food or water. The first problem may be that the intended poison makes the bait unacceptable to animals that have never encountered the poison. This is called primary bait refusal. Because of this common difficulty, many efforts to find better rodenticides have emphasized highly toxic substances of such bland taste and odor that animals always will take a lethal dose the first time. However, this is an impractical objective. At least a few animals will get only a sublethal dose on first encounter and will be conditioned thereby to avoid the poison even though it seems tasteless and odorless at first. This reaction is called secondary bait refusal or bait shyness. There is even some indication that rodents learn from the behavior of their companions in such a way that the manner of death of some of them conditions the behavior of others that have consumed no poison. A third problem with many rodenticides is that they are very nearly as dangerous to humans and useful animals as to rodents. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

This problem can be minimized by selecting a poison with a wide margin of safety, by coloring the bait, by combining the poison with an emetic, or by restricting the placement of baits. However, all these solutions have their limitations. There is no rat poison that cannot harm humans if sufficiently misused. Considering all these difficulties, four requirements for an ideal rodenticide may be stated as follows: (a) the poison must be surely effective when incorporated into baits in such small quantity that its presence is not detected to an interfering degree. (b) Finished baits containing the poison must not excite bait shyness in any way and the necessity of prebaiting must thereby be avoided. (c) The manner of death must be such that surviving individuals will not become suspicious of its cause but will remain on the premises and eat freely of the bait until they themselves die. (d) The poison, in the concentration used for control, must be specific for the species to be destroyed unless its use can be made safe for humans and domestic animals by some other means. Part of the safety of the anticoagulant rodenticides is made possible by their cumulative properties and depends on the fact that they are offered to rodents in such a way that a single dose is harmless even to the rodents themselves. Quite aside from the important species differences in susceptibility, which favor human safety, people are protected further by the fact that except in suicide or murder, substantial continuing exposure is far less likely than a single accidental exposure. Although this chapter is devoted to synthetic organic rodenticides, it is necessary to recall that inorganic and botanical compounds may still be important for rodent control in some areas. Furthermore, some of them, notably arsenic, phosphorus, and strychnine, are very important as sources of human poisoning. On the contrary, some synthetic rodenticides have other uses. The most important examples are the use of vitamin D and of certain anticoagulants as rodenticides and as drugs in human medicine. In addition, several of the organic fluorine compounds have been used experimentally or in practice as systemic insecticides and/or raticides. 2153

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100.2.1  Sodium Fluoroacetate 100.2.1.1  Identity, Properties, and Uses (a) Chemical Name Sodium monofluoroacetate is the chemical name. (b) Structure See Figure 100.1. (c) Synonyms Sodium monofluoroacetate is also known as Compound1080 or ten-eighty. The CAS registry number is 62-74-8. (d) Physical and Chemical Properties The empirical formula for sodium monofluoroacetate is C2H2FNaO2 and the molecular weight is 100.3. Its forms

Sodium monofluoroacetate

FCH2— C – NH2 2-Fluoroacetamide O

– –

Sodium fluoroacetate came to prominence in the United States as a result of a search for rodenticides that would not be subject to shortages imposed by World War II (Ward, 1945). This and related compounds had been considered earlier as systemic insecticides. At about the same time, it became known that fluoroacetate is the toxic material in the South African plant gifblaar (Dichapetalum cymosum). Later it was shown that the same compound is present intermittently in Acacia georgiana. The main toxicant in Dichapetalum toxicarium is fluorooleic acid, but fluoropalmitic acid is present also. The ground seeds of D. toxicarium have been used by natives as a rat poison. It gave problems with secondary poisoning and human toxicity similar to that later associated with synthetic sodium fluoroacetate. Under these circumstances, there were practical as well as academic reasons to study the mode of action of organic monofluoro compounds. It appears that the toxicity of all of the compounds depends on the same mechanism. Highly toxic compounds either have two carbon atoms or are metabolized to this form (Chenoweth, 1949; Peters, 1963b; Raasch, 1958; Saunders, 1947).

O

– –

100.2  Fluoroacetic acid and its derivatives

O FCH2—C–ONa – –

FCH2—CH2–OH

FCH2 — C – N – CH3

2-Fluoroethanol MNFA OH

–

OH

–

Compared to toxic substances in general, biochemical actions of the synthetic rodenticides that have been studied in humans are unusually well known. This makes it possible to assign them to groups (see the following sections) that are meaningful not only chemically but in terms of biochemical lesions. The same is not true of numerous miscellaneous compounds including crimidine (BSI, ISO) that apparently have not yet produced poisoning in humans and certainly have not been used as drugs or studied experi­ mentally in human subjects.

FCH2 — CH2 – CH2F

FCH2 — C–— CH2Cl

Glycerol difluorohydrin

Glycerol chlorofluorohydrin

Figure 100.1  Some organic fluorine rodenticides and other organic fluorine pesticides.

an odorless, white, nonvolatile powder that decomposes at about 200°C. Although the compound is often said to be tasteless, dilute solutions actually tasted like weak vinegar. Sodium fluoroacetate is very water-soluble and hygroscopic but is of low solubility in ethanol, acetone, and petroleum oils. (e) Formulations and Uses Sodium fluoroacetate is formulated as an aqueous solution containing a warning color. Sodium monofluoroacetate is used to kill rats, mice, other rodents, and predators. It is an intense mammalian poison, and it is used in several countries, essentially in New Zealand where it is used on a large scale to control the introduced Australian brush-tail possum (Trichosurus vulpecula), which is the main source of contamination of cattle and deer herds by bovine tuberculosis. It is also used to a lesser extent in Australia. It is used only by trained personnel.

100.2.1.2  Toxicity to Laboratory Animals (a) Basic Findings The first paper on sodium monofluoroacetate as a rodenticide (Kalmbach, 1945) drew attention to its very high acute toxicity. LD50 values for ordinary laboratory rats and for wild animals of the same species were reported as 2.5 and 5.0 mg/kg, respectively. The wild black rat (Rattus rattus), another commensal species, was much more susceptible (LD50: 0.1 mg/kg). An LD50 of 0.22 mg/kg has been reported for Rattus norvegicus (Dieke and Richter, 1946). The likelihood of danger to people, domestic animals, pets, and nontarget wildlife was pointed out. The acute toxicity of the compound to an extremely wide range of wildlife was reported by Ward and Spencer (1947) (Table 100.1). Fluoroacetate acts mainly on the central nervous system and the heart. It seems that there are species in which fluoro­acetate affects chiefly the heart, such as the rabbit, the goat, and the horse, and others in which only the central nervous system is affected, such as the dog, the guinea pig,

Chapter | 100  Rodenticides

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Table 100.1  Single-Dose LD50 for Sodium Fluoroacetate Species

Route

LD50 (mg/kg)

Reference

Rat

Oral

0.22

Dieke and Richter (1946)

Rat

Oral

2.5

Ward and Spencer (1947)

Rat

Oral

1–2

Phillips and Worden (1957)

Rat

Intraperitoneal

3–5

Ward and Spencer (1947)

Mouse

Subcutaneous

19.3

Hutchens et al. (1949)

Mouse

Subcutaneous

17.0

Tourtellotte and Coon (1951)

Mouse

Intraperitoneal

10.0

Ward and Spencer (1947)

Mouse

Intraperitoneal

16.5

Tourtellotte and Coon (1951)

Mouse

Intraperitoneal

14.7

Raasch (1958)

Guinea pig

Oral

0.4

Ward and Spencer (1947)

Guinea pig

Intraperitoneal

0.37

Hutchens et al. (1949)

Rabbit

Subcutaneous

0.28

Hutchens et al. (1949)

Dog

Oral

0.06

Tourtellotte and Coon (1951)

Cow

Oral

0.39

Robinson (1970)

Calf

Oral

0.22

Robinson (1970)

Opossum

Oral

0.79

Bell (1972)

Mallard duck

Oral

4.8

Hudson et al. (1972)

South African clawed toad (Xenopus laevis)

Oral

500

Chenoweth (1949)

and the frog. In the cat, the rhesus monkey, the domestic pig, and birds, both systems are involved. The above results were obtained by Chenoweth and Gilman (1946) using methyl fluoroacetate instead of sodium fluoroacetate. However, since both compounds yield the fluorocitrate ion in the body, where it is converted to fluoroacetate, which is responsible for the induction of pharmacologic and toxic signs (see below), it seems that this experiment is nevertheless interesting in showing a large degree of species variability in the site of action. In all species, there was a delay of 0.5–2 h or more between administration, either oral or intravenous, and the onset of the symptoms, and the route of administration did not significantly affect the toxicity of fluoroacetate. Laboratory rats acquire a tolerance to sodium fluoro­ acetate by ingesting sublethal doses over a period of 5–14 days. However, this tolerance is lost if intake of the compound is interrupted for as little as 7 days (Kalmbach, 1945). Tolerance of some but not all species was confirmed by several investigators, including Kandel and Chenoweth (1952). These authors found that, whereas small doses of fluoroacetate increased tolerance to challenge doses of fluoroacetate or 4-fluorobutyrate, tolerance to neither could be evoked by small doses of 4-fluorobutyrate. The citrate content of the rat brain appeared to have no relation to tolerance, and the citrate that accumulated after a small dose

did not prevent the further accumulation of citrate after a larger dose. The sensitivity of mice to sodium fluoroacetate depends on temperature. Under otherwise identical conditions, the LD50 values were 12.1 and 5.16 mg/kg at 23 and 17°C, respectively (Misustova et al., 1969). The survival of individual rats in a particular dosage group may be predicted by following their body temperature. There is a critical level that varies somewhat according to the interval after dosing. Animals that regained their initial temperature within 96 h usually lived, but those that failed to regain normal temperature within this time usually died (Filip et al., 1970). For groups of animals, the course of the temperature can be described by a computer-generated curve (Hosek and Love, 1952). The temperature change is correlated with citrate metabolism (Kirzon et al., 1970). The lagtime between when monofluoroacetate is consumed and the occurrence of the first symptoms of poisoning in mammals is from 0.5 to 3 h. Clinical signs of acute poisoning include rapid and labored breathing, tremors and muscle spasms, terminal convulsions, and death. It was shown that the cause of death in herbivores is cardiac failure, whereas in carnivores it is central nervous system disturbances with convulsions and respiratory failure (Egekeze and Oehme, 1979). Toxic effects induced by subchronic daily oral administration of sodium monofluoroacetate to Sprague-Dawley

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rats of both sexes for 90 days have been reported by Eason and Turck (2002). Microscopic changes were restricted to the testes and heart of male rats exposed to the highest dose (0.25 mg/kg bw/day) and included severe hypospermia in the epididymes, severe degeneration of the seminiferous tubules, and cardiomyopathy. The no-observed-effect level (NOEL) was the intermediate dose level of 0.075 mg/kg bw/day. Primates and birds are more resistant; rodents and carnivores are most susceptible. In general, cold-blooded vertebrates are less sensitive than warm-blooded ones (Egekeze and Oehme, 1979). Sodium monofluoroacetate in carcasses creates a secondary poisoning hazard to which carnivorous predators are extremely susceptible (Bell, 1972). Pathological changes resulting from poisoning in sheep have been reported. They include scattered foci of fibrous tissue in the myocardium and mild degeneration of the striatum and hippocampus in the brain. Plasma elimination half-life in rabbits was shown to be 1.1 h and the retention time in tissue greater with larger doses. Tissues residues were substantially lower than in plasma (Gooneratne et al., 1995). When orally administered to sheep and goats at dose levels of 0.1 mg/kg body weight, the plasma elimination half-life was found to be 10.8 h in sheep and 5.4 h in goats. Concentrations of sodium fluoroacetate in muscle, kidney, and liver (0.042, 0.057, and 0.021 g/g, respectively) were clearly lower than those in the plasma (0.098 g/g) 2.5 h after administration. After 96 h, only traces of the compound were detectable in sheep tissues (0.002–0.008 g/g). The authors concluded that even with accidental exposure to sublethal doses, sodium fluoroacetate would not persist in tissues for more than a few days because of its rapid clearance and because occurrence of residues in meat intended for human consumption would be highly unlikely (Eason et al., 1994). (b) Absorption, Distribution, Metabolism, and Excretion Using ether as a solvent, it was possible to recover 60–70% of the total dose from the body (including gastrointestinal contents) of rabbits killed by 10 times the LD50 level. The concentration in the brain was twice that in other organs (Tomiya et al., 1976). Sodium monofluoroacetate is rapidly absorbed by the gastrointestinal tract. It is not well absorbed by the intact skin, but absorption may be greater in the presence of dermatitis or other skin injury (Eason, 2002). Studies in mice have shown that 1080 concentrations in plasma, muscle, and liver decrease by half after less than two following oral administration. Prolonged persistence of the compound in animal tissues after sublethal oral admininistration is unlikely as shown in rabbits, goats, possums, and sheep (Eason et al., 1994). (c) Biochemical Effects It was in connection with the mode of action of fluoroacetic acid that the term “lethal synthesis” was coined

Hayes’ Handbook of Pesticide Toxicology

(Peters, 1952). Peters (1963a) later reviewed the research and extended the concept. Very briefly, no mammalian enzyme was found that was inhibited by fluoroacetate in vitro. However, in vivo, the ion undergoes synthesis to form fluorocitrate and this inhibits mitochondrial aconitase either in vivo or in vitro. The result is that the Krebs cycle is blocked, which leads to lowered energy production, reduced oxygen consumption, and reduced cellular concentration of ATP; furthermore, since the citrate synthetase continues to work, citrate accumulates in the tissues (Buffa and Peters, 1950). It is thought that toxicity is due not to the accumulation of citrate per se but to the blockage of energy metabolism. However, increased tissue and plasma concentration of citrate is probably responsible for some of the symptoms seen during acute poisoning. Citrate is a potent chelator of calcium ion, and it has been demonstrated that in cats intravenously injected with fluoroacetate at 0.03 mmol/kg the ionized calcium level in blood fell by an average of 27.2%, 40 min after the injection. There was a corresponding prolongation of the QT interval of the electrocardiogram (ECG), and treatment with CaCl2 significantly prolonged the life of the treated animals as compared with unmedicated positive controls (Roy et al., 1980). The characteristic delay at the onset of poisoning by sodium fluoroacetate is accounted for by the time necessary for its metabolism and biochemical mode of action. The toxicity of fluoroacetate is entirely different from that of inorganic fluorides. It depends on the firmness of the F—C bond such that fluoroacetate is an antimetabolite. Consistent with the theory that fluorocitrate is the active toxicant, it was found to be at least 100 times more toxic than fluoroacetate when injected directly into the brain under various experimental conditions. An intracerebral dose of 0.115 g failed to kill rats weighing about 250 g, and it did not cause convulsions; doses of 0.287 g (about 0.001 mg/kg) or greater caused convulsions and killed almost all rats (Morselli et al., 1968). On the other hand, a dosage of 40–60 mg/kg is necessary to kill by the intraperitoneal route, and an oral dosage of 40 mg/kg constitutes only an LD50. The great difference was attributed to failure of fluorocitrate to reach aconitase within critical cells of the brain and heart (Peters and Shorthouse, 1971). Species differ in the degree to which the concentration of citrate increases in different organs and also in the timing of these increases (Kirzon et al., 1973). These biochemical differences presumably underlie the clinical differences between species, especially the relative importance in neurological and cardiac effects. Accumulation of citrate was evident in mice within 2 h after intraperitoneal injection of sodium fluoroacetate at a rate of 30 mg/kg, which is about 1.7 times the LD50 in that species. The concentration of citrate increased from 48 ppm in controls to 74, 101, and 166 ppm within 2, 5, and 24 h, respectively, after injection. The mice were dead at 24 h (Matsumura and O’Brien, 1963).

Chapter | 100  Rodenticides

Whereas Williamson et al. (1964) agreed that the initial effect of fluoroacetate is to produce fluorocitrate, they considered that the secondary inhibition of phosphofructokinase by the accumulated citrate was actually lethal because it deprived the cell of pyruvate, which would eventually overcome the inhibition of aconitase. (d) Effects on Organs and Tissues Loracher and Lux (1974) concluded on the basis of studies of neuromembrane depolarization that diminished inhibitory conductance is apparently important as a causative factor in convulsions induced by sodium fluoroacetate. The decreased level of ionized calcium in blood induced by the chelating effect of citrate certainly plays a role in the depolarization of the neuromembrane, as it does on the cardiac cell membranes. The effect of sodium fluoroacetate on the heart rhythm is due, as demonstrated by Noguchi et al. (1966), primarily to action on the cells themselves and not on the vagus nerve. Irregularity of rhythm and a condition analogous to fibrillation were produced in cultures of heart cells that had grown until cell-to-cell contact was prevalent and beating was synchronized. The average times necessary to produce irregularity and fibrillation were 9 and 48 h, respectively, at a concentration of 10 ppm in the medium, but only 2 and 9 h, respectively, at a concentration of 100 ppm. At a concentration of 1000 ppm, fibrillation was immediate and cytoplasmic vacuoles appeared rapidly. (e) Effects on Reproduction A single dosage of sodium fluoroacetate just below the maternal LD50 reduced oxygen consumption of the embryos as well as the mother but was not teratogenic (Spielmann et al., 1973). However, with relatively high doses (0.33 and 0.75 mg/kg bw/day) for about one-third of the gestation duration, limited skeletal effects were noted and the NOEL for developmental effects was 0.1 mg/kg bw/day considering the occurrence of abnormal ribs at 0.33 mg/kg bw/day (Eason et al., 1999). (f) Treatment of Poisoning in Animals Hutchens et al. (1949) demonstrated a significant reduction of mortality in mice, guinea pigs, and rabbits (but not dogs) treated with ethanol at a rate of 800 mg/kg administered subcutaneously as a 10% solution in normal saline. The response occurred when the alcohol was given before signs of poisoning appeared and was best when given within 10 min of poisoning. In mice, sodium acetate and ethanol acted synergistically to antagonize poisoning (Tourtellotte and Coon, 1951). The beneficial effect of ethanol in rodents was confirmed by Chenoweth et al. (1951), but these authors found ethanol less effective in the dog and utterly useless in the monkey. In a study of a wide range of chemical substances in mice, rats, rabbits,

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dogs, and rhesus monkeys, they concluded that commercially available monacetin containing about 60% glycerol monoacetate was superior to any other substance tested as an antidote for poisoning by fluoroacetate. Not only did it reduce mortality, but it was able to normalize heart and brain rhythms as indicated by ECG and electroencephalogram (EEG) tracings. Light pentobarbital anesthesia for 18–24 h significantly reduced mortality among dogs poisoned by sodium monofluoroacetate at a rate of 0.10 mg/kg (Hutchens et al., 1949; Tourtellotte and Coon, 1951).

100.2.1.3  Toxicity to Humans (a) Accidental and Intentional Poisoning Sodium fluoroacetate was introduced in 1946 in the United States for use by pest control operators, including persons hired for the purpose by government agencies. The poison was mixed with a dye. Solutions were supposed to be placed in shallow paper cups made in such a way that they would not tip over. These water baits were supposed to be used only in places that would be unoccupied and locked during exposure of the poison, and all cups and dead rodents were supposed to be collected and incinerated by authorized persons at the end of the exposure period. However, the regulations were not always followed. By the end of the year, at least one child who found an “empty” paper cup had died, and her 3-year-old brother had been severely poisoned. By the end of 1949, there had been at least 12 deaths and six cases of nonfatal poisoning. In addition, there had been four deaths, all in children, that probably were caused by sodium monofluoro-­ acetate, but other sources of poisoning could not be ruled out. Of the 12 deaths clearly caused by sodium monofluoro­ acetate, five involved small children who had found and often chewed on a poison cup, three involved juveniles who had found the poison in a soft drink bottle, and four were suicides of adults. Except one, each of the survivors was a child who had found a poison cup. These accidents made such an impression on the few people who had legal access to sodium fluoroacetate that they became far stricter in carrying out the recommended precautions and in selecting situations in which the compound was used at all. As a result, the safety record of the compound in the United States improved greatly. A typical fatal case involved a 40-year-old man who was found unconscious in his bedroom. He had an 8-year history of severe depression, and his family had been warned of the possibility of suicide. When admitted to the hospital, he had slight muscular spasms and nystagmus of both eyes; the heart rate was 92 beats per minute and rhythm was irregular. Following gastric lavage and a soft soap enema, the nystagmus became worse, and the patient had an epileptiform convulsion. The blood pressure fell to 90/40 mmHg. Treatment consisted of plasma, oxygen, and

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procaine hydrochloride in the hope of desensitizing the heart. The blood pressure improved to 118/75 mgHg, but there was no decisive change until the heart and later the respiration stopped about 17 h after admission (Harrisson et al., 1952b). Another fatal case was remarkable for its combination of prolonged survival following the ingestion of an almost certainly very large dose. Briefly, about 113,000 mg of sodium fluoroacetate was missing from a professional rat exterminator’s supplies after his 17-yearold son made a solution and drank it. The boy vomited promptly and then within 1 h walked into an hospital emergency room. He gradually became comatose during gastric lavage, and consciousness was never regained. Within less than 3 h of ingestion, he had a grand mal convulsion associated with fecal incontinence. The clinical course, which lasted slightly over 5 days, was characterized by cardiac irregularity, which responded to a considerable degree to procainamide hydrochloride; dilation and failure of the heart with acute pulmonary edema, which responded surprisingly well to digitalis (lanatoside C); bouts of severe hypotension, which responded only questionably to levarterenol (norepinephrine) but somewhat better to mephentermine; cortical irritability, which responded to barbiturates and later responded more effectively to ethanol; frequent severe carpopedal spasm, controlled somewhat by calcium gluconate; and finally growing evidence of infection including a temperature reaching 42.3°C in spite of efforts to reduce it. The diagnosis based on autopsy was poisoning, bronchopneumonia with septicemia, focal infarction of the right kidney, and mediastinal emphysema (Brockmann et al., 1955). Serious illness followed by full recovery occurred in a 2-year-old boy who was found licking crystals from the screw cap of a bottle of sodium fluoroacetate solution. The parents did not know whether he drank any of the solution. Almost immediately after he was found, the boy began to vomit. He was brought to hospital about 6 h later because he began to have generalized convulsive movements and became stuporous. On admission, the boy was comatose and exhibiting carpopedal spasms, tetanic convulsive movements, irregular respiration, and great cardiac irregularity. While a solution of calcium gluconate was being injected, there were a few seconds of cardiac asystole. Thereafter, the irregular cardiac rhythm resumed but at a much slower rate. Tetanic convulsions stopped immediately, and the child became completely flaccid. A few hours after admission, the child became responsive. Very soon the boy suffered a generalized tonic clonic convulsion lasting several minutes and followed by deep coma. Briefly, the boy remained unresponsive for 4 days. Cardiac rhythm continued to change frequently during the first 3 days. Tonic convulsions lasting several minutes occurred many times every hour, sometimes about every 10 min for many successive hours. During spasm, the pupils dilated and remained inactive to light; between seizures the pupils

Hayes’ Handbook of Pesticide Toxicology

were miotic but responsive to light. On two occasions respiration stopped and artificial respiration was required briefly. On the evening of the 4th day, 100 h after ingestion, the boy began to open his eyes and look about. He tried to talk but was unable to articulate. He could neither sit up nor reach for objects but appeared alert. On the 5th and 6th days, he rapidly regained all his motor ability, slowly lost his drowsiness, and became articulate. On the evening of the 6th day he was clinically well. He was discharged on the 11th day. Reexamination 1 year later showed that the boy had had no further neurological trouble, and his mental and physical development has proceeded normally (Gajdusek and Luther, 1950). In another case in which the initial dosage undoubtedly was smaller, there were no important clinical changes until 20 h after ingestion, when the 8-month-old girl had a generalized seizure lasting about 1 min. In spite of treatment with phenobarbital, three additional seizures occurred during the next 12 h. There was no further illness, and the patient was discharged 4 days later. Follow-ups revealed no change in behavior, intellect, or motor performance (Reigart et al., 1975). Any serious but reversible interference with respiration or general circulation is liable to produce some cases in which the patient survives but with severe brain damage. The cardiac arrhythmias characteristic of poisoning by sodium fluoroacetate are likely to produce such interference. An example involved an 8-year-old boy who was in status epilepticus when he entered hospital. The convulsions were controlled to some degree. There was no striking change until 14 h after admission, when ventricular asystole occurred. Heart action was renewed but only after sufficient delay that the child suffered brain damage and was clearly mentally defective after a very long and stormy hospital course (McTaggart, 1970). During the decade 1971–1981, 111 cases of accidental or unintentional poisoning with sodium fluoroacetate were collected by the National Poison Center of Israel. These cases included three cases of death and one case of mass accidental poisoning affecting 30 children, although the great majority of them only consumed a very small number of wheat grain baits impregnated with the compound. These latter cases did not result in clinical symptoms of poisoning (Roy et al., 1982). These authors also described the clinical features of two cases of acute poisoning in which gastrointestinal disorders were rapidly followed by central nervous system manifestations (disorders of consciousness, convulsions, coma) and cardiac disorders, the most frequent cause of death. Ventricular ectopic beats preceded the ventricular arrhythmia, which was then followed by ventricular tachycardia and fibrillation. The electrocardiogram was characterized by a prolonged QT interval. A metabolic acidosis was commonly observed. Chung (1984) reported on five cases collected between 1975 and 1981

Chapter | 100  Rodenticides

in Taiwan. The amount ingested ranged from 8 to 40 ml of a 1% formulation of sodium monofluoroacetate. All five patients survived. All cases had signs of transient cardiac dysfunction, but in addition acute renal failure was seen in three of the five patients, two of them with frank uremia. The acute renal failure was reversible. In a retrospective study of 38 cases of poisoning collected between 1988 and 1993 in Taiwan by Chi et al. (1996), 18% of the patients died. Laboratory symptoms included nonspecific ST-T and T waves on the ECG (72%), hypocalcemia (42%), and hypokalemia (65%). Hypotension, respiratory rate, pulse rate, increased serum creatinine, and decreased pH were considered as the most important predictors of mortality. The recently published case of a 15-year-old girl’s attempted suicide is interesting because of the unusual pathological sequelae. About 1 h after absorbing an unknown quantity of sodium monofluoroacetate, she developed a grand mal seizure associated with tachycardia (150/min), followed by three additional seizures over the ensuing 4 h and became comatose. She recovered but developed a chronic cerebellar ataxia and computerized tomography findings of moderate diffuse brain atrophy (Anonymous, 2007a). (b) Use Experience In spite of the great toxicity of sodium fluoroacetate, there apparently has been only one case of illness among those who used it without suicidal intent. Even in this case, the kind of illness was so atypical of poisoning and so complicated by the unrelated factor of hypertrophy of the prostate that evaluation of the case is difficult. The patient entered hospital with renal failure and other serious illness. Although he was only 59 years old, he had a 5-year history of symptoms of prostatism but no history of urinary tract infection, renal calculi, or hematuria. He had had gout for 10 years, and he had been digitalized for 12 months. For 6 months he had experienced increasing lassitude, vomiting, and pruritus. Inspection revealed rapid breathing and muscle wasting. More detailed physical examination revealed mild left ventricular failure and evidence of liver disease, hypothyroidism, extrapyramidal disease, and gout, as well as distended bladder, caused by prostatic hypertrophy. These findings were substantiated by laboratory examinations. Following catheter drainage of the bladder, blood urea declined and renal function improved further following prostatectomy 10 days after admission. Recovery was very slow. Neurological and thyroid findings cleared within 6 months. Renal function continued to improve for about 2 years, after which the patient remained well. Involvement of fluoroacetate was suspected because of the history of exposure, the finding of organic fluorine in the urine, and histological changes found in kidney biopsies. There was no doubt of exposure; the patient had been employed for 10 years as an exterminator of rabbits, and

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for about 4 weeks each year he had applied sodium fluoro­ acetate to pieces of carrots that served as bait for the animals. During this work, he had worn rubber gloves and he had never knowingly ingested any of the poison. The report of concentrations of 15.4 and 14.8 ppm of sodium fluoro­ acetate (analyzed as organic fluoride) in two samples of urine collected 2 weeks after admission and the absence of such organic fluoride in samples collected 5 weeks and 6 months later was accepted as consistent with the history of exposure. A kidney biopsy performed 4 days after admission revealed periglomerular fibrosis, some capsular adhesion and other glomerular changes, plus swelling and vacuolation of tubular cells, increased interstitial fibrous tissue, a few small foci of inflammation, and mild thickening of the arterial walls. A second biopsy 4 weeks later showed little change in the glomerular lesion, but the tubules were no longer vacuolated. However, many tubules had been lost, and many of those remaining were atrophic. Interstitial fibrosis was prominent. The kidney lesions were considered similar to those described in rats in association with acute poisoning. It was acknowledged that lower urinary tract obstruction may have been a predisposing factor (Parkin et al., 1977). Even if the patient had been exposed to sodium fluoroacetate a short time before he was admitted to hospital, it is difficult to understand why excretion of organic fluoride from this source would continue 2 weeks later. Although no urinary levels of organic fluoride have been reported for other workers, one must note that 15 ppm would indicate a daily output of about 22.5 mg/person/day in a person with average urinary volume. This in turn would indicate a minimal absorption rate of about 0.32 mg/kg/day, an astonishingly high level. The renal changes previously described in rats (Cater and Peters, 1961) followed one or a few very large dosages of fluorocitrate, and the fat droplets were tiny compared to those seen in the human patient. (c) Dosage Response In a fatal case, 465 mg (equivalent to a dosage of over 6 mg/kg) was recovered from the stomach contents, urine, brain, liver, and kidneys (Harrisson et al., 1952a). No account was taken of sodium fluoroacetate in other organs and tissues or of that removed by vomiting, lavage, and enema; therefore, the ingested dosage must have been considerably larger. Several children varying in age from 0.66 to 8.0 years were poisoned seriously or even fatally by chewing on only one paper cup placed earlier for rat control. The cups were made to receive 15 ml of 0.33% solution, that is, 50 mg of sodium fluoroacetate. The average age of the children was 2.37 years, and the weight of such a child is about 13 kg. Thus, the maximal dosage must have been approximately 3.8 mg/kg, but the true dosage must have been considerably smaller because part of the material originally added to the cup may have been lost and not all that dried in the cup would have been ingested. A dosage of 0.5–2.0 mg/kg must

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be considered highly dangerous. The estimated mean lethal dose in humans ranged from 2 to 10 mg/kg (Gajdusek and Luther, 1950; Harrisson et al., 1952b). (d) Laboratory Findings The following concentrations expressed as sodium fluoroacetate were found in samples taken at autopsy from a man who survived about 17 h after being found unconscious: urine, 368 ppm; liver, 58 ppm; brain, 76 ppm; and kidney, 65 ppm (Harrisson et al., 1952a). Recently, an HPLC with fluorescence detection analytical method in biological samples (serum, food, meat) has been described (Xie et al., 2007). (e) Pathology In a fatal case, autopsy revealed petechial hemorrhages and congestion of the organs consistent with recent fits. All the findings were nonspecific, but it is interesting that they included diffuse tubular degeneration of the kidneys, which is consistent with the findings in the only case of alleged chronic human poisoning by sodium fluoroacetate. (f) Treatment of Poisoning Apparently, most patients who survived poisoning by sodium fluoroacetate as well as those who died of it received no medication that offered any possibility of specific antidotal action. In at least one case (unpublished), a poisoned child was treated with whiskey and survived. Unfortunately, no details are available, and there can be no assurance that the child would not have progressed equally well without treatment. Although monacetin apparently has not been administered to a human patient, the work of Chenoweth et al. (1951) in various animals, especially monkeys, offered good reason to think it would be valuable for treating human poisoning. They recommended that it be injected intramuscularly at least every hour for several hours at the rate of 0.1–0.5 ml/kg per injection. There is no clinical evidence for or against the use of acetate in humans. On the contrary, acetamide has been administered to patients, and it seemed to be the reason for their survival. It is available at Accident and Emergency Departments throughout New Zealand. Acetamide is administered intravenously as a 10% solution in 5% glucose. In severe cases, 500 ml is given in 30 min every 4 h; in milder cases, 200 ml is given on the same schedule. There can be no doubt that removal of the poison and supportive care are indicated. A number of patients have shown clear-cut poisoning but survived without sequelae following such treatment. Supportive care should include continuous cardiac monitoring. There is strong clinical evidence that the danger of cardiac arrhythmia can be reduced significantly by judicious and continuing use of procainamide hydrochloride. Even so, equipment for defibrillation should be ready. There is reason to hope it would be successful if required because at least one

patient was revived with only external massage of the heart. There is also clinical evidence that cortical irritability can be lessened by barbiturates. There is no basis for speculating on the value of diazepam in this connection. Contrary to the evidence in monkeys, clinical evidence in humans has indicated that ethanol is beneficial and perhaps superior to barbiturates. Whereas the effect seemed to involve needed sedation, the possibility of a more fundamental effect in the biochemical lesion was not excluded. On the basis of laboratory studies, Chenoweth et al. (1951) recommended against administration of calcium, potassium, sodium chloride, bicarbonate, or acetate. They considered that any necessary replacement of fluid should be done cautiously with plasma, and they considered digitalization as definitely contraindicated. However, clinical experience argues strongly against two of these prohibitions, and there is no clinical evidence to support some of the others. Calcium gluconate has proved useful in controlling carpopedal spasm, including such spasm in a patient who survived without sequelae. Digitalis (lanatoside C) not only improved the function of a poisoned heart that had failed to the point of acute pulmonary edema, but also produced no detectable side effects. Finally, there is clinical evidence that mephentermine is more effective than levarterenol in raising blood pressure if that becomes necessary in the course of poisoning by sodium fluoroacetate.

100.2.2  Fluoroacetamide 100.2.2.1  Identity, Properties, and Uses (a) Chemical Name 2-Fluoroacetamide is the chemical name. (b) Structure See Figure 100.1. (c) Synonyms Fluoroacetamide is also known as Compound 1081. Trade names for fluoroacetamide include Fuorakil, Fussol, Megarox, and Yancock. The CAS registry number is 640-19-7. (d) Physical and Chemical Properties Fluoroacetamide has the empirical formula C2H4FNO and a molecular weight of 77.06. It is a crystalline solid that sublimes on heating but melts at 107–109°C. It is very soluble in water, moderately soluble in acetone, and sparingly soluble in aliphatic and aromatic hydrocarbons. (e) History, Formulations, and Uses At one time fluoroacetamide was used as a systemic insecticide for scale insects, aphids, and mites on fruits; however, it has been considered too toxic to mammals for commercial

Chapter | 100  Rodenticides

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use as an insecticide. Its use as a rodenticide was suggested by Chapman and Phillips in 1955. It is used as a bait (20 g active ingredient/kg) in areas to which the public have no access, such as sewers and locked warehouses. It is formulated as dyed cereal-based bait that is mixed with water for use.

100.2.2.2  Toxicity in Laboratory Animals (a) Basic Findings Fluoroacetamide is a compound of moderate to high acute toxicity depending on the species (see Table 100.2). In the WHO Recommended Classification of Pesticides by Hazards (World Health Organization, 1986), the technical material is listed in class IB, “Highly hazardous.” The compound is absorbed by the skin (Phillips and Worden, 1957). Animals acutely poisoned by this compound show listlessness, irritability, chronic convulsions, abasia, piloerection, and irregular respiration (Araki, 1972). One characteristic usually observed in animals dying from acute poisoning with fluoroacetamide as well as with sodium fluoroacetate is postmortem rigidity (Bentley and Greaves, 1960). Death generally occurs in coma after convulsions have stopped (Phillips and Worden, 1957). There is no obvious difference in susceptibility between the sexes (Bentley and Greaves, 1960). The time that elapses between

Table 100.2  Single-Dose LD50 for Fluoroacetamide Species

Route

LD50 (mg/kg)

Reference

Rat

Oral

15

Phillips and Worden (1957)

Rat

Oral

13

Bentley and Greaves (1960)

Rat

Dermal

20a

Phillips and Worden (1957)

Mouse

Oral

30.62

Araki (1972)

Mouse

Subcutaneous

34.20

Araki (1972)

Mouse

Intraperitoneal

85

Matsumura and O’Brien (1963)

Rabbit

Oral

1.5–2.0

Phillips and Worden (1957)

Rabbit

Intravenous

0.25

Buckle et al. (1949)

Chicken

Oral

4.25

Egyed and Shlosberg (1977)

a

Lowest lethal dose.

dosing and the onset of convulsions appears to be related to the dosage level, and fluoroacetamide affects behavior (Bentley and Greaves, 1960). Subacutely poisoned animals show anorexia, emaciation, and alopecia (Araki, 1972). Perhaps because of strain differences, investigations have reported slightly different thresholds for the largest repeated dosage tolerated by rats without clinical signs. As discussed later, the threshold for testicular injury is much lower. Phillips and Worden (1957) found that 3 mg/kg/day for 20 days was without effect on appetite or general health. Mazzanti et al. (1964) found similar results in rats on a dietary level of 50 ppm (about 2.5 mg/kg/day for 90 days). However, Steinberger and Sud (1970) reported that this same dietary level caused a reduction of food intake and of growth. The poisoning of farm animals by effluent from a factory that manufactured fluoroacetamide caused the Ministry of Agriculture, Fisheries, and Food to recommend that the compound should not be used as an insecticide in agriculture, for home gardens, or food storage in Great Britain, and it was withdrawn from the market (Allcroft and Jones, 1969; Allcroft et al., 1969; Anonymous, 1964a,b). (b) Absorption, Distribution, Metabolism, and Excretion Investigators agree that fluoroacetamide is less toxic than fluoroacetate. This has been attributed to the fact that metabolism of the former compared to the latter is slower (Matsumura and O’Brien, 1963). In fact, Phillips and Worden (1957) reported that they recovered, from the urine of rats receiving fluoroacetamide at a rate of 3 mg/kg/day, 62% of the total intake unmetabolized, and they confirmed the identity of the compound by melting point and mixed melting point. This finding raises the possibility that the toxicity of fluoroacetamide (albeit lower than that of fluoro­ acetate) is in part inherent and does not depend entirely on metabolism to fluoroacetamide in the rat testis, an effect apparently not reported for fluoroacetate. (c) Biochemical Effects Evidence that fluoroacetamide has essentially the same mode of action as sodium fluoroacetate is the finding that mammals poisoned by the amide contain greatly elevated levels of citrate (Allcroft et al., 1969; Egyed and Brisk, 1965; Egyed and Miller, 1971; Egyed and Shlosberg, 1977; Matsumura and O’Brien, 1963). Further evidence is offered by the fact that cockroaches convert fluoroacetamide to fluoroacetate as well as to fluorocitrate, and mouse amidase hydrolyzes fluoro­ acetamide (Matsumura and O’Brien, 1963). (d) Effects on Reproduction Selective destruction of the germinal epithelium of the testes of male rats apparently was reported first by Mazzanti et al. (1964), who studied only a single dosage level resulting from a dietary level of 50 ppm. On this diet, the body

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weight of 150- to 160-g rats increased by 88% in 90 days but the testes were reduced to slightly less than one-third of the weight in controls. After 64 days, the tubules were almost completely lacking in seminal cells; only some spermatogonia, the Sertoli cells, and the interstitial cells were apparently undamaged. Peculiar giant cells were observed. It was noted that fluoroacetamide acts first on the more mature cells of the germinal epithelium and not on the cells where mitoses are more numerous. Dividing cells in the intestinal mucosa were undamaged. In a later study, male rats that received a dietary level of 50 ppm (usually calculated as about 2.5 mg/kg/day but said to be about 3.4 mg/kg/day in these rats) showed a marked morphological change in the nucleus of tep-13 spermatids within 24 h, and the effects became more pronounced and the entire cell became distorted in 5 days. After 10 days of treatment, earlier-step spermatids showed degenerative changes and giant cell formation. Eventually, even spermatocytes were affected. Androgen secretion by the testis apparently was not affected. Dietary levels of 20, 10, and 5 ppm produced characteristic changes in late-stage spermatids but no effect on spermatocytes. The 5-ppm level had no effect on the weight of the testis of rats fed as long as 28 days, but higher levels led to a marked decrease in weight. Subcutaneous administration of fluoro­ acetamide at a rate of about 1.0 mg/kg/day produced the characteristic change in stage-13 spermatids within 4 days and a 50% reduction in the weight of the testis in 28 days. Spermatogenesis continued, and spermatocytes and young spermatids remained apparently normal, but late spermatids were distinctly abnormal. Subcutaneous doses of about 0.2, 0.04, and 0.02 mg/kg/day produced little or no change in the weight of the testis and produced progressively less histological injury so that change was barely discernible at the lowest dosages. Thus, the effect of fluoroacetamide on spermatogenesis is specific and not secondary to general toxicity, which (in the form of reduced growth) was evident only at the highest oral dosage (Steinberger and Sud, 1970). Testicular degeneration caused by fluoroacetamide has been confirmed in rats and reported in other species (Egyed, 1973). Fluoroacetamide at an oral dosage of 15 mg/kg also interferes with reproduction in female mice, whether administered 2 days before or 10 days after fertilization; pregnancy was prolonged, prenatal mortality was increased, and the young suffered from cyanosis, respiratory distress, reduced growth, and decreased survival (Tokavera et al., 1971). (e) Effects on Wildlife and Nontarget Species In some countries, fluoroacetamide is used to control field rodents, thus exposing nontarget species to either direct toxic effects by feeding on the baits or secondary effects by feeding on carcasses of rodents killed by the compound. These effects have been experimentally studied by Braverman (1979) on several nontarget species such

Hayes’ Handbook of Pesticide Toxicology

as mongoose (Herpestes ichneumon), hyena (Hyaena hyaena), snakes, birds, cats, and dogs. This experiment showed a degree of susceptibility of the animals similar to that reported for sodium fluoroacetate; it also confirmed that the dog was the most sensitive species. Some species showed a relative tolerance to direct poisoning; this was the case for barn owls, buzzards, and the black kite. A secondary poisoning study was done by offering the carnivore carcasses of birds (Meriones tristrami) that had fed freely on poisoned grains. The results were quite variable; the mongoose was the most susceptible, whereas the risk of secondary poisoning to birds of prey was not high. An outbreak of poisoning by fluoroacetamide in four greylag geese (Anser anser) and teal (Anas crecca) has been reported in Israel (Shlosberg et al., 1975). Clinical signs in one goose were described as severe convulsions, incoordinated twisting of the neck, total anemia, prostrating depression, and death. (f) Treatment of Poisoning in Animals Sodium acetate did not protect rats poisoned by fluoroacetamide. However, when administered as a mixture by mouth at a ratio of 4:1 or 9:1, acetamide raised the LD50 of fluoroacetamide from 15 to 22 mg/kg. When acetamide was administered by mouth at a dosage of 180 mg/kg within 65 min or less after fluoroacetamide at the otherwise fatal oral dosage of 20 mg/kg, all rats survived. The same was true when the ratio (9:1) remained the same, the delay did not exceed 60 min, and the dosage of poison was as high as 35 mg/kg. However, the antidote was ineffective when the delay was 105 min or greater (Phillips and Worden, 1956). When administered to rats as a mixture by mouth, l-cysteine hydrochloride was antidotal, raising the LD50 of fluoro­ acetamide from 15 to 25 and 30 mg/kg, respectively, at dosage ratios of 4:1 and 9:1 (Phillips and Worden, 1957). Acetamide at an oral dosage of 2500 mg/kg was also effective in treating chickens when given within 20 min after fluoroacetamide at a dosage of 10 mg/kg (slightly more than twice the LD50 level). The same dosage given 30 min after the poison or 500 mg/kg given with the poison were ineffective (Egyed and Shlosberg, 1977). In limited tests, neither acetamide nor monoacetin was effective in treating poisoned sheep (Egyed, 1971). The ineffectiveness of sodium acetate and apparently of monoacetin and the effectiveness of acetamide and l-cysteine as antidotes for poisoning by fluoroacetamide raise the possibility that the effective compounds do not prevent biochemical lesions directly but rather competitively retard the conversion of fluoroacetamide to fluoroacetate and thus permit more time for excretion of unmetabolized fluoroacetamide.

100.2.2.3  Toxicity to Humans (a) Accidental and Intentional Poisoning At about 11:30 h, an 18-month-old girl removed a 120-ml bottle of 1% fluoroacetamide from a low drawer in the family

Chapter | 100  Rodenticides

kitchen and drank some of the contents. On the advice of a pharmacist, the child was given olive oil, the white of an egg, and milk at about noon and was made slightly sick. The child remained lively and played in the garden until her usual bedtime, 18:30. A about 23:30 that evening the child vomited but was put back to bed when she appeared all right. Apparently the child was not checked until 10:30 h next morning, when she was found in a semiconscious state. On a physician’s order, she was taken to hospital, but convulsions occurred on the way and the patient arrived about 11:30 h in a shocked state. The child was given about 10 ml acetamide in water once, 3.7 ml of brandy in water each hour, and symptomatic treatment. She continued to have occasional convulsions and remained unconscious until she died almost 96 h after ingesting the poison. Both the heart and kidney contained 6.3 mg of organic fluoride per gram of dry tissue; the citrate content (108 ppm in heart and 23.9 ppm in kidney) was not considered significantly high. From the evidence available, it was estimated that the baby had consumed about 300 mg of fluoroacetamide or 23 mg/kg (Great Britain Ministry of Agriculture, Fisheries, and Food, 1961; WHO, 1963). An analytical method was developed for fast determination of fluoroacetamide in body fluids. It is a GC/MS method after solid-phase microextraction (SPME), with acetamide as an internal standard (Cai et al., 2004). (b) Treatment of Poisoning Treatment of poisoning by fluoroacetamide should be the same as that for fluoroacetate (see Section 100.2.1.3) with due attention to removal of the poison and general care of the patient. Based on animal studies, rapid and energetic treatment with acetamide is recommended. A dosage of 315 mg/kg was effective in rats, but a much higher dosage was required in chickens. It is of special importance that the first dose be given at the earliest possible moment. Repeated administration was not used in the animal experiments but would appear wise. A combination of intravenous monoacetin (glyceryl monoacetate, 0.55 gm/kg), sodium acetate (0.12 gm/kg), and ethanol (0.12 gm/kg) has also been recommended (Dipalma, 1981). Dipalma also suggested, as an alternative course, the oral administration of 100 ml of monoacetin plus 500 ml of water every hour for about 2 h. Hemoperfusion involving fixed-bed uncoated charcoal was used in one case, but it was not helpful and the patient died (de Torrente et al., 1979).

100.2.3  Fluoroethanol 100.2.3.1  Identity, Properties, and Uses (a) Chemical Name 2-Fluoroethanol is the chemical name.

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(b) Structure See Figure 100.1. (c) Physical and Chemical Properties Fluoroethanol has the empirical formula C2H5OF and a molecular weight of 64.07. It is a solid melting at approximately room temperature (26.5°C). It has a density of 1.091, a boiling point of 103°C and a flash point of 31°C. (d) Use It is used as a rodenticide.

100.2.3.2  Toxicity to Laboratory Animals Fluoroethanol is a compound of high acute toxicity, as indicated by an intraperitoneal LD50 of 5 mg/kg in the rat (Bartlett, 1952). According to Bartlett, fluoroethanol is relatively inactive and its toxicity depends on its oxidation to fluoroacetate by tissue alcohol dehydrogenase.

100.2.3.3  Toxicity to Humans Three cases of poisoning of workers by fluoroethanol occurred in a chemical plant, in at least two instances, as the result of accidental rupture of a container and rapid evaporation of the fluid. A typical patient suffered onset in about 90 min and was discharged from hospital in 4 days. All patients had tremor, severe muscular weakness, nausea, headache, and a slight swelling of the liver. (Hemorrhagic gingivitis in one patient and prediabetic hyperglycemia in another were explained by their past histories and were unrelated to poisoning.) Examination of the other 40 workers in the plant failed to reveal any complaints or clinical finding that could be related to the compound (Colamussi et al., 1970). There is no specific treatment for subacute poisoning except, of course, complete cessation of exposure. If acute poisoning should occur, it should be treated like poisoning by fluoroacetate (see Section 100.2.1.3). This compound does not seem to be marketed any longer.

100.3  Substituted ureas One of the compounds that has been promoted as a rodenticide relatively safe for other mammals is pyriminil, a substituted urea (see Figure 100.2). It is not clear whether this group of compounds has been explored extensively with a view to selecting the one with the best combination of effectiveness for killing rodents and safety for humans and useful animals. However, it has become apparent that pyriminil and some other substituted ureas are specific poisons for the  cells on the pancreas and, therefore, cause diabetes mellitus. This effect may not be related to the mode of action of pyrimidil as a rodenticide, but it has great bearing on the overall safety of the material.

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2164

– –

O

S

–

– –

NH—— C —— NH

NH —C —NH2

CH2

N

NO2 Pyriminil O

–

CH3 —N—C—NH ON

ANTU H

OH H H

O OH

HO

HN

C

NH O

O

H H

O

CH2 — OH Streptozocin

Alloxan

Figure 100.2  ANTU, a thiourea rodenticide, and three substituted ureas known to cause diabetes in one or more species.

100.3.1  Pyriminil

Table 100.3  Single-Dose LD50 for Pyrimidil in Various Species Species

Sex

Oral LD50 (mg/kg)

Albino rat

M

12.30

Norway rat

M

4.75

Roof rat

M

18.00

Cotton rat

M, F

20–60

Albino mouse

M

84

House mouse

M

98

Deer mouse

M

10–20

Guinea pig

M

30–100

Rabbit

M

300

Dog

M

500

Cat

M, F

62

Rhesus monkey

M, F

2000–4000

Pig

M

500

Vole

M, F

205

Chicken

M

710

Pigeon

M, F

1780

100.3.1.1  Identity, Properties, and Uses (a) Chemical Name N-(3-pyridylmethyl)-N’-(4-nitrophenyl)-urea is the chemical name. (b) Structure See Figure 100.2. (c) Synonyms Pyriminil is also known as PNU, pyrinuron, and RH-787. It was sold under the trade names Vacor, Rat Killer, DLP787 20% bait, and DLP-787 10% House Mouse Tracking Powder. The CAS registry number is 53558-25-1. (d) Physical and Chemical Properties Pyriminil has the empirical formula C13H12N4O3 and a molecular weight of 272.27. It decomposes at 223°C. (e) History, Formulations, and Uses Pyriminil was introduced in 1975 and developed as an acute rodenticide. It was used to control Norway rats, roof rats, and house mice; it was especially effective against rodents resistant to anticoagulant poisons. Pyriminil was sold for indoor use only as a prepared bait containing 2% active ingredient and a 10% tracking powder. The product was withdrawn from the market by the U.S. manufacturer in 1979 (Chappelka, 1980), but it is still manufactured on a small scale for local use in the People’s Republic of China, for example.

100.3.1.2  Toxicity in Laboratory Animals There are greater differences in the susceptibility of different species to pyriminil (technical material) as shown in Table 100.3 (Peardon, 1974). The marked susceptibility of Norway rats was of course the basis for its use as a rodenticide. Cats also are very susceptible. Apparently, a good description of the signs of acute poisoning in laboratory animals has not been published. A simple list of signs and symptoms in dogs has been given in the distribution company’s technical bulletin. The onset of the symptoms may be delayed 4–48 h. They include nausea and emesis, depression, initial constriction of pupils followed later by dilated pupils and visual impairment with slow pupillary response to light, ataxia, fine to coarse tremors, hind-limb weakness, decreased reflexes, deep breathing, and dehydration. Similar symptoms have been reported in a horse that had eaten at least 250,000 mg (about 250 mg/kg). The animal showed severe muscular fasciculations, dilated pupils, and profuse sweating within 24 h after ingestion. Laboratory tests revealed severe hyperglycemia (418 mg/100 ml) and indications of liver injury (elevated liver enzymes). The animal was treated with intravenous nicotinic acid (2.2 mg/kg) followed by four subsequent injections of 1 g and recovered; it was considered clinically normal 3 months later. Three other poisoned horses showed the same signs as well as intense abdominal pain, hind-limb weakness, ataxia, and

Chapter | 100  Rodenticides

persistent inappetence (Russell et al., 1978). Peoples and Maddy (1979) have reported, without details, poisoning in domestic animals (two horses, three cats, and 17 dogs) in California. The case of a 22-kg dog seen eating a full 30g packet of Vacor (780 mg active ingredient) is mentioned. Immediately following ingestion, the dog vomited but became blind 2 days later. (a) Absorption, Distribution, Metabolism, and Excretion Pyriminil is rapidly absorbed by rats, mice, and dogs after oral administration. Blood levels peaked in 1–6 h, depending on species and site of the radiolabel 14C. Gastrointestinal transit of 14C is more rapid in dogs than in rats. Urinary and fecal excretions are of similar importance in all three species. Tissue distribution of two 14C labels (nitrophenyl and pyridyl) varied, especially in dogs. The liver contained more of the dose than any other single organ (Deckert et al., 1978a). Rats tolerated, metabolized, and eliminated single or multiple sublethal dosages (5 mg/kg) but were less efficient than dogs in detoxifying dosages in excess of 20 mg/kg. It was concluded that the tolerance of dogs for the compound depended on their efficient hepatic extraction, metabolism, and excretion of it (Deckert et al., 1978b). Several metabolites of pyriminil have been identified (Deckert et al., 1978a, 1979). These include aminopyriminil, p-aminophenyl urea, p-acetamidophenyl urea, p-nitroaniline, p-phenylenediamine, p-acetamidoaniline, nicotinic acid, nicotinuric acid, and nicotinamide. The concentrations of these metabolites varied from one species to another. The presence of the parent compound in rat and human urine suggests that they may be more sensitive to the compound than the dog because of less efficient metabolism (Deckert et al., 1979). (b) Biochemical Effects Repeated, sublethal doses of pyriminil increased the urinary and fecal excretion of a later dose of the compound tagged with 14C; however, the same animals showed increased hexobarbital sleeping time and other evidence of inhibition of certain liver microsomal enzymes, especially p-nitroanisole O-demethylase. Whatever microsomal enzymes are responsible for metabolism of pyriminil are induced by pretreatment with 3-methylcholanthrene, which increases the biliary excretion of the metabolites and decreases pyriminil toxicity 50-fold (Deckert et al., 1977, 1978a). Mild pyriminil-induced hyperglycemia was observed in rats; it was also shown to be reversible by insulin (Deckert et al., 1977). The diabetogenic effects of pyriminil were also confirmed in patients poisoned by the product. This effect is the result of a direct toxic action of the  cells of the pancreas. Wilson and Gaines (1983) have demonstrated that pyriminil at concentrations ranging from 102 to 105 M preferentially intoxicates rat pancreatic  cells

2165

in culture, within 1 h of contact. It was also shown in the study that nicotinamide can reduce pyriminil-induced  cell injury, thus confirming previous findings by Karam et al. (1980) that nicotinamide could partially reverse pyriminil inhibition of glucose-stimulated insulin secretion by freshly isolated islets of Langerhans from the rat. In addition to its diabetogenic effect, pyriminil has a direct effect on glucose metabolism. The erythrocytes of patients poisoned by the compound showed a marked depression of glucose consumption as well as decreased uptake of methylene blue in the presence of glucose. In addition, a 0.1-mM concentration in vitro caused decreased utilization of glucose and decreased uptake of methylene blue by erythrocytes from normal people and rabbits (Lee and Lee, 1977). The mechanism of action has been investigated and it was shown that pyriminil specifically inhibits the NADH: ubiquinone reductase activity of complex I in mammalian mitochondria. The activity of other respiratory enzymes of mitochondria is unaffected at concentrations that completely inhibit the redox and energetic function of complex I. Inhibition of complex I activity quantitatively correlates with the inhibition of insulin release in insulinoma cells and pancreatic islets and is also consistent with the doses reported in cases of human poisoning. These results indicate that the toxic and diabetogenic action of pyriminil primarily derives from the inhibition of mitochondrial respiration of NAD-linked substrates in the high-energy demanding pancreatic islets (Esposi et al., 1996). (c) Treatment of Poisoning in Animals The mechanism of action of pyriminil remains uncertain, but it is of interest that alloxan, streptozotocin, and dithizone, all which can induce diabetes mellitus in experimental animals, are substituted ureas. However, it would appear that some species such as dogs, cats, and laboratory primates are refractory to both the diabetogenic and neurotoxic effects of pyriminil (Karam et al., 1980). Whereas 6-aminonicotinamide is not a substitute urea, it is toxic to  cells and it is a recognized antagonist of nicotinamide (Herken, 1971). Because nicotinamide can prevent the toxic effect of streptozotocin (Ganda et al., 1976), alloxan (Rossini et al., 1975), and N-3-pyridylmethl-N4nitrophenyl urea (Deckert et al., 1977), it seems possible that all of these compounds act as nicotinamide antagonists.

100.3.1.3  Toxicity to Humans (a) Accidental and Intentional Poisoning There are many reports on human poisoning in the literature describing the main clinical and laboratory features of those poisonings. A 25-year-old man with a history of psychiatric disturbances attempted suicide by injecting an unknown amount of pulverized methaqualone tablets and ingesting two packets of rat poison, each containing 737 mg of

2166

pyriminil. Seven days later he was admitted to a local hospital for treatment of a staphylococcal abscess of the left antecubital fossa. He received antibiotics and had rapid clinical improvement. It was recorded that since attempting suicide the patient had noticed lassitude, anorexia, abdominal bloating, constipation, and the onset of painful paresthesia with numbness of his legs and difficulty in walking. A random plasma glucose level on admission was 309 mg/100 ml, and check samples taken on subsequent days were slightly higher. Ketones and glucose were present in the urine. On the 4th hospital day, insulin therapy was started. The diabetes gradually was controlled, although tolerance for carbohydrate and need for insulin were erratic. An upper gastrointestinal tract series done on the 10th hospital day showed gastric and proximal smallbowel hypomobility bordering on atony. The patient was discharged on hospital day 19 on a regimen of insulin and temporary thoridiazine. The patient remained well for 16 days and then returned because of nausea and vomiting. He was found to have severe autonomic and peripheral polyneuropathy characterized by orthostatic hypotension, greatly diminished response to pinprick and vibratory sensation in the lower extremities, and other changes. Although the diabetes was now better controlled, the serum sodium was low (116 mEq/l), and the syndrome of inappropriate antidiuretic hormone was demonstrated. The hyponatremia responded to fluid restriction, and the orthostatic hypotension was improved by support stockings. Ten months after the suicide attempt, the patient experienced two episodes of weakness and lethargy that were relieved by eating. He had lost about 18 kg and appeared cachectic (45 kg, 174 cm) but alert and well oriented. His gait was ataxic and there was substantial muscle wasting. A very thorough examination showed reduced disappearance rate of intravenous glucose and depressed C-peptide response to intravenous glucose when compared with a normal control but no impairment of glucagon release after stimulation by intravenous arginine. Nerve conduction studies demonstrated severe sensory and mild motor neuropathy. Quadricep capillary basement membrane thickness was in the diabetic range. Insulin was discontinued and tolbutamide prescribed. Following discharge, the patient regained 5 kg and experienced subjective improvement of his neuropathy (Prosser and Karam, 1978). Whereas most clinical studies have placed greatest emphasis on the diabetogenic action of pyriminil, its injury to the nervous system was no less remarkable, as emphasized in a paper by LeWitt (1980a). This injury often involved autonomic impairment (postural hypotension often severe enough to cause fainting when the patient sat up, impaired pupillary responses, impotence, decreased sweating, urinary retention, dysphagia, and gastrointestinal hypomobility), peripheral neuropathy (loss of musclestretch reflexes, sensory loss, neurogenic myopathy), and

Hayes’ Handbook of Pesticide Toxicology

encephalopathic and dyskinetic features (loss of cortical function ranging from confusion to coma, cerebellar ataxia, tremor, motor hyperactivity, nystagmus, and diffuse electroencephalographic changes). In addition, some cases involved chest or epigastric pain and some showed ischemic electrocardiographic changes. Cardiac arrhythmias were occasionally the cause of death. Neurological disorders often appeared within hours after ingestion. Occasionally, onset was delayed or insidious. Symptoms related to different parts of the nervous system began and later improved at different times in the same patient, and the order of progression varied from case to case. Neurological improvement took many months, and full recovery was uncommon, orthostatic hypotension in particular tending to persist. Causes of delayed death included inanition, sepsis, aspiration pneumonia, and insulininduced hypoglycemia. Accidental ingestion of pyriminil by a 25-month-old boy resulted in acute vomiting, lethargy, seizures, hypoglycemia (followed by hyperglycemia and glucose intolerance), and autonomic and peripheral neuropathy (Johnson et al., 1980). A review of reports unpublished in 1978 indicated seven deaths and two nonfatal cases in Korea and four fatal and 11 nonfatal cases in the United States. At least in the United States, all the cases were in adults; all but one were attempted suicide; all the survivors developed diabetes mellitus and autonomic nervous system dysfunction, chiefly dysphasia, dystonia, and bowel and bladder dysfunction. Hypothermia and paresthesias were seen. A later review revealed nearly 90 cases in the United States and over 250 in Korea (Frethold et al., 1980). A case of acute poisoning (approximately 67 mg/kg) in a 42-year-old man with all the signs already described but characterized by a severe orthostatic hypotension with full spontaneous recovery 11 months after hospitalization was reported by Osterman et al. (1981). Gallanosa et al. (1981) have compared the main features of four cases reported with enough details in the literature with those of one case of their own. (b) Dosage Response A dose as low as 780 mg was fatal within 150 days. A dose of 2340 mg was fatal within 1 day, but a patient survived 40 days after ingesting 7020 mg. One patient survived 2340 mg, and at least two survived 1560 mg but not without characteristic, persistent illness. The smallest dose known to have produced characteristic illness was 390 mg (about 5.6 mg/kg) (LeWitt, 1980b). (c) Laboratory Findings The most important findings for guiding treatment and often for diagnosis include nearly transient hypoglycemia followed by persistent hyperglycemia, glycosuria, ketosis, and elevation of serum amylase and lipase activities. p-Nitroaniline

Chapter | 100  Rodenticides

at a concentration of 5.1 ppm has been reported in the liver of a person who died after accidentally ingesting pyriminil (Osteryoung et al., 1977). In the case of a 7-year-old boy who was found dead a day after another child saw him ingest a packet of pyriminil, unchanged compound at a concentration of 1.5 ppm was found in the urine hydrolysate and two metabolites were found in the liver and some other samples. Aminopyriminil (nitro group metabolized to an amine) was found at concentrations of 5.6, 1.4, 0.3, and 0.6 ppm in liver, kidney, spleen, and urine, respectively. Acetamidopyriminil (amino group conjugated with acetic acid) was found in traces in the blood and liver (Frethold et al., 1980). Karam et al. (1980) reported (in addition to the clinical features) autopsy findings from several cases of acute poisoning, including that of a 7-year-old boy. All three cases showed extensive islet degeneration of the pancreatic tissue with generalized destruction of  cells and sparing of  and  cells as well as of the exocrine glandular tissue. Islet-cell surface antibodies were detected in four of the six reported cases. It may be that these antibodies are the result rather than the cause of -cell destruction. (c) Pathology Loss of  cells of the pancreas has been observed generally in persons killed by pyriminil (Frethold et al., 1980; Karam et al., 1980; LeWitt, 1980a; Prosser and Karam, 1978). Lesions of the nervous system have not been found so regularly. In one case reported by LeWitt (1980a), no lesions of the central or peripheral nervous system were found; in another case, cerebral edema and neuropathic changes restricted to the sensory spinal roots were found. Autopsy of a 39-year-old man who survived 19 days revealed (a) severe loss of ganglion cells and rare degenerating neurons in the paravertebral sympathetic ganglia, (b) marked loss of neurons in the sensory spinal ganglia with multiple residual nodules of Nageotte, (c) marked degeneration of the sensory roots and posterior columns, (d) slight perivascular lymphatic infiltrates in both the sympathetic and sensory ganglia, (e) swelling of nerve fibers and thinning of the myelin sheaths of the sural nerve, and (f) isolated degenerated and regenerating fibers in the skeletal muscles (Papasozomenos, 1980). (d) Treatment of Poisoning Patients who develop diabetes mellitus clearly must be treated for that condition in the usual way. There is good reason from animal experiments to believe that diabetes could be prevented if the patient were given large, repeated doses of nicotinamide beginning promptly after ingestion of the poison. However, cases have ended in diabetes and neuropathy when nicotinamide was started 9 and 14 h, respectively, after ingestion. Nicotinamide was considered possibly beneficial in the case of an infant, even though administration was started something over 12 h after ingestion of pyriminil (Johnson et al., 1980). However, the fact

2167

that the child received the poison “on a piece of gum” offered by another child suggests that the initial dose was small, and complete recovery may have been due to that fact alone (Pont et al., 1979). The dose and duration of the treatment with nicotinamide are still uncertain (Anonymous, 1979). Nicotinic acid has also been tried as an antidote (Pont et al., 1979), but its use is contraindicated because (a) it is toxic in humans, (b) it protects animals only against alloxan and not streptozotocin (Ganda et al., 1976), and (c) its vasodilatory effects may complicate the control of blood pressure. Cases that require insulin may progress so that insulin is no longer required, but the patient can be maintained on sulfonylureas.

100.4  Thioureas The development of ANTU as a poison for adult Norway rats was described by Richter (1945). The entire development was a result of a chance observation associated with studying the taste of phenylthiourea, which is bitter to most people but tasteless to a few who inherit this specific lack of sensation as a Mendelian recessive trait. When an attempt was made to explore this taste difference in animals, it was found that if a few crystals were placed on the tongues of rats, all of them died overnight. The wide and prolonged use of phenylthiourea for taste and inheritance tests without any untoward effect indicated its safety for humans, whereas the results in rats suggested that it might serve as a rat poison. Further study revealed that rats detect and reject phenylthiourea too effectively for it to be practical as poison. This led to a systematic search of other thiourea derivatives with high toxicity but little or no taste. All monosubstituted thiourea derivatives tested produced pulmonary edema and pleural effusion in the laboratory rat (Dieke et al., 1947). The toxicity of thiourea to wild Norway rats was enhanced by a single aromatic substitution on one of the nitrogen atoms. Two or more substitutions on one or both nitrogen atoms lowered the toxicity, as was also true of substitution on the sulfur atom. ANTU was chosen as the most suitable compound. Although the dog is susceptible to ANTU, most animals, including monkeys, are resistant. This offered the hope that humans would be resistant also, and extensive field trials in areas of Baltimore led to no toxic symptoms either in workers or in the over 500,000 persons living in the treated areas (Richter, 1945). A disadvantage of ANTU as a rodenticide is that young Norway rats and roof rats of all ages are too resistant to the compound for it to be practical for their control. Another disadvantage is the prompt appearance of both tolerance and bait refusal in adult Norway rats that have received a nonfatal dose. Tolerance is completely lost within 30 days,

2168

but refusal may last longer (Richter (1945, 1946). Gaines and Hayes (1952) found that bait shyness lasted at least 4 months under field conditions. Several interesting observations were made during the survey of thioureas. All of these compounds produce hyperplasia of the thyroid gland. Whereas nonlethal doses of unsubstituted thiourea have little effect on pigmentation or hair growth, phenylthiourea destroys pigment both in the skin and in the hair but without affecting growth of the hair, and ANTU completely stops pigment production and growth of hair. Withdrawal of the substituted thioureas is followed in less than 10 days by recovery of pigment and hair growth. Finally, different strains of Norway rats on different diets showed thiourea LD50 values as different as 4 and 1830 mg/kg. The difference was modified but not eliminated by placing the rats on the same diet as that of the most susceptible ones. Age differences in the susceptibility of Norway rats to thiourea are similar to those with ANTU (Dieke and Richter, 1945; Richter, 1945).

100.4.1  ANTU 100.4.1.1  Identity, Properties, and Uses (a) Chemical name 1-(1-Naphthyl)-2-thiourea is the chemical name. (b) Structure See Figure 100.2. (c) Synonyms ANTU, an acronym for -naphthylthiourea, is the approved common name (BSI, ISO) for this compound. Trade names include Anturat, Bantu, Kill Kantz, Krysid, Rattrak, and Rat-tu. Code designations for ANTU include Chemical-109 and U-5227. The CAS registry number is 86-88-4. (d) Physical and Chemical Properties ANTU has the empirical formula C11H10N2S and a molecular weight of 202.27. Pure ANTU forms colorless crystals and the technical grade is a gray crystalline powder with a bitter taste. Its melting point is 198°C (pure). Its solubility in water at 25°C is 0.06 g/100 ml; in acetone, 2.43 g/100 ml; and in triethylene glycol, 8.6 g/100 ml (technical). (e) History, Formulations, and Uses ANTU was discovered as a rodenticide in 1945. The formulations include baits (10–30 g/kg) and tracking powders (200 g/kg). It is used specifically against the Norway rat. In some countries it has been withdrawn from use because of the carcinogenicity of -naphthylamines present as impurities (Worthing and Walker, 1983).

Hayes’ Handbook of Pesticide Toxicology

100.4.1.2  Toxicity in Laboratory Animals (a) Basic Findings Different investigators have been in good agreement about the acute oral toxicity of ANTU to Norway rats. Dieke and Richter (1945) and Lehman (1951, 1952) found oral LD50 values of 6.9 and 6 mg/kg, respectively. There is a wide variation in susceptibility among different species, especially to intraperitoneal administration (see Table 100.4). Rats and dogs are the most susceptible species and rabbits the least susceptible one with an oral LD50 of 1000 mg/kg bw (McClosky and Smith, 1945). The Norway rat is particularly susceptible, the young being slightly more resistant than the adults. (b) Absorption, Distribution, Metabolism, and Excretion Early studies on the metabolism of phenylthiourea and of diphenylthiourea suggest the basis of the toxicity of monosubstituted thioureas. It was shown by Dieke et al. (1947) that the oral LD50 of the phenyl compound is 8.6 mg/kg, whereas a dosage of 2000 mg/kg of the diphenyl compound did not produce illness. Both rats and rabbits excrete little phenylthiourea as compounds with the —C  S group intact (Carroll and Noble, 1949; Williams, 1959). In rabbits, the proportion of such compounds was only about 12%, but the corresponding proportion was about 70–80% for diphenylthiourea. These observations suggest that toxicity was associated with desulfuration in vivo (Williams, 1959). It has been speculated that ANTU acts on the lung by the release of hydrogen sulfide (Petit et al., 1970), but this seems highly unlikely because rats rendered tolerant to ANTU are not tolerant to hydrogen sulfide (or carboxyl sulfide or phosgene) (Carroll and Noble, 1949). (c) Biochemical Effects By using a mixture of 35S- and 14C-labeled ANTU, it was possible to show that some of the sulfur and a smaller proportion of the carbon were covalently bound to macromolecules of the lung and liver following in vivo administration. By contrast, practically no radioactive carbon was bound when an equal amount of the almost nontoxic, 14C-labeled oxygen analogue of ANTU (-[14C]naphthylurea) was administered. In the presence of NADPH, ANTU was metabolized by either lung or liver microsomes in vitro in such a way that the rates of binding of 35S or 14C to macromolecules of the microsomes were greater than those associated with boiled microsomes or with normal microsomes without NADPH. Binding in the presence of active enzyme and NADPH was covalent and accompanied by a decrease in the level of cytochrome P450 detectable as its carbon monoxide complex. Pretreatment of rats at the rate of 2 mg/kg/day for 5 days produced a decrease of their microsomal enzyme activity as measured by metabolism of parathion. All such pretreated rats survived a

Chapter | 100  Rodenticides

2169

Table 100.4  Single-Dose LD50 for ANTUa Species

Route

LD50 (mg/ kg)

Reference

Rat

Intraperitoneal

10

Boyd and Neal (1976)

Rat

Intraperitoneal

7

Rat

Intraperitoneal

5

Norway, domestic I

Intraperitoneal

2.5

Norway, domestic II

Intraperitoneal

6.25

Norway, wild, adult

Intraperitoneal

6.20–8.10

Norway, wild, young

Intraperitoneal

16–58

Alexandrine

Intraperitoneal

250

Norway, wild

Oral

6.9

Mouse

Intraperitoneal

56

Rabbit

Intraperitoneal

400

Guinea pig

Intraperitoneal

140

Guinea pig

Intraperitoneal

350

Dog

Intraperitoneal

16

Dog

Dermal

38

Cat

Oral

500

Monkey

Intraperitoneal

175

Monkey

Oral

4250

Chicken

Intraperitoneal

2500

Chicken

Oral

4250

DuBois et al. (1947)

a

From IARC (1983).

dosage of ANTU (10 mg/kg) which killed six to 10 controls, and binding of 35S by proteins of the liver and especially the lungs of the pretreated animals was less than that of the controls. Pretreatment with 4-ipomeanol protected all rats from an otherwise uniformly fatal dose of [35S]ANTU and caused a slight reduction of covalent binding of 35S to lung (but not liver) proteins. Finally, rats pretreated with dimethylmaleate, which depletes tissue stores of glutathione, were killed by ANTU at 5 mg/kg, a dosage which was harmless to controls. In every instance, rats killed by ANTU showed a hydrothorax of at least 4 ml, whereas those protected by pretreatment with ANTU or ipomeanol developed no hydrothorax. These findings were interpreted as evidence that (a) the toxicity of ANTU

depends on metabolic activation and on covalent binding of the reactant(s) to lung macromolecules and (b) tolerance to ANTU is the result of inhibition of microsomal enzymes and consequent reduction in the metabolic activation of a challenge dose (Boyd and Neal, 1976). An extension of this reasoning would attribute the normal tolerance of young Norway rats to ANTU to their relative lack of microsomal enzyme activity. Further study showed that about half of the atomic sulfur released from ANTU reacted with cysteine side chains of microsomal protein to form a hydrodisulfide. The other moiety released by microsomal enzymes is -naphthylurea (Lee et al., 1960). (d) Effects on Organs and Tissues ANTU induced reverse mutations in Salmonella typhimurium strain TA1538 in the presence but not in the absence of Arocolor- or phenobarbital-induced rat liver microsomal preparations. A preparation purified by thinlayer chromatography was as active as a technical grade material, thus excluding the attribution of activity to impurities. ANTU also transformed Syrian hamster embryo cells in vitro without the addition of an activating system (Kawalek et al., 1979). The International Agency for Research on Cancer (IARC) has determined that ANTU was mutagenic to bacteria but did not induce unscheduled DNA synthesis in rat hepatocytes in vitro (IARC, 1987). ANTU was tested for carcinogenicity (Fitzburg and Nelson, 1947) in mice (Innes et al., 1969) by administration in the diet. No tumor was reported in either study, but the International Agency for Research of Cancer (IARC, 1983) found that both studies were inadequate to evaluate the carcinogenicity of ANTU to experimental animals. It was concluded that ANTU was not classifiable as to its carcinogenicity to humans (Group 3). (e) Pathology As far as the rat lung is concerned, ANTU causes marked edema of the subepithelial spaces of the alveolar walls without erosion or other damage to type I and type II epithelial cells. Thus, edema caused by ANTU differs morphologically and presumably in mechanism from that produced by injection of epinephrine or by an injection of a mixture of fibrinogen and thrombin into the cerebrospinal cistern (Hatakeyama and Shigei, 1971). The edema caused by intraperitoneal ANTU in rats is dosage-related in the range of 3–50 mg/kg. Although interstitial edema was the first observable change, bleeding and scalloping of endothelial cells were observed within 2 h, and epithelial damage was apparent electron microscopically within 6 h following 50 mg/kg. The injury was apparently similar to but more rapid than that caused by 99% oxygen at 1 atmosphere pressure (Meyrick et al., 1972). Not only pulmonary

2170

edema but also pleural effusion shows a dosage–response relationship (Sobonya and Kleinerman, 1973). Using a different approach, Bohm (1973) demonstrated changes which he interpreted as indicating increased permeability to colloidal carbon in the pulmonary arteriodes as well as capillaries and venules of rats within 3.25 h after an intraperitoneal injection of ANTU at a rate of 10 mg/kg. In anesthetized sheep given 20, 50, 75, or 100 mg/kg ANTU intravenously, the first phase of the response consisted of transient increases in pulmonary artery pressure and plasma and lymph thromboxane B2 concentrations. These changes were not dependent on the dose of ANTU administered. At 2–4 h after administration; pulmonary artery pressure and thromboxane concentrations were normal or near normal. ANTU produces a two-phase response with the steady state characterized by a dose-dependent increase in lung microvascular permeability (Havill et al., 1982). These authors, on the basis of experimental results in sheep, suggest that the severe pulmonary hypertension that follows ANTU administration may be mediated by vasoconstrictor products of arachidonic acid metabolism and that the complement or coagulation systems may be involved as well, resulting in pulmonary microemboli. O’Brien et al. (1985) have reported that isolated lungs from rats treated 4 h earlier with ANTU had decreased conversion of angiotensin I to angiotensin II and that the extent of decrease was related to the dose of ANTU administered and to the perfusate flow rate. It may be that permeability of membranes of the kidney as well as those of the lung and pleura is increased inasmuch as urinary excretion of albumin occurs (Patil and Radhakrishnamurty, 1977). Fantone et al. (1984) have investigated the effects of prostaglandin E1 (PGE1) on ANTU-induced lung injury and have shown that systemic administration of a stable analogue of PGE1 potentiated those effects in a dose- and time-related manner when subcutaneously administered 1 h prior to ANTU treatment (1 mg/kg bw intraperitoneally). Increased lung and pleural effusion was associated with increased pulmonary endothelial cell bleeding and gap formation with a decrease in the number of platelet thrombi in the PEG1-treated animals. Those results have not yet been confirmed. It was experimentally shown that intratracheal pretreatment of SPF Sprague-Dawley rats with keratinocyte growth factor significantly reduced the acute permeability pulmonary edema in -naphthylthiourea. Histopathology showed abundant type II pneumocyte hyperplasia in lungs of animals pretreated with keratinocyte growth factor and marked pulmonary edema in animals pretreated with naphthylthiourea. Less edema was observed in the keratinocyte growth factor/-naphthylthiourea-treated group. The authors have suggested that the likely mechanism of this protection was related to type II pneumocyte hyperplasia (Mason et al., 1996).

Hayes’ Handbook of Pesticide Toxicology

In a subsequent study, Guery et al. (1997) have confirmed that keratinocyte growth factor, through type II alveolar pneumocyte hyperplasia with increased sodiumpotassium-adenosine triphosphatase activity, attenuated ANTU-induced edema formation by potentiating alveolar fluid clearance. The mechanism of action responsible for the known resistance to -naphthylthiourea effects in certain circumstances has been investigated. It was noted during the development phase of this rodenticide that rats become resistant to the lethal effects of ANTU if they are first exposed to a small, nonlethal dose of ANTU. It was also observed that young rats were resistant. In this study, groups of male Sprague-Dawley rats were orally treated with a small nonlethal dose (5 mg/animal) 24 h before challenge with a 100% lethal dose (70 mg/animal). The protection induced by the low dose of ANTU lasted for 5 days and slowly decreased, totally disappearing by day 20 postadministration. The high dose of ANTU alone resulted in an increase in pulmonary edema leading to the animals’ death. In the low-dose rats, a stimulation of pulmonary cell hyperplasia was observed. Treatment with the antimitotic agent colchicine abolished the low-dose ANTU-induced resistance against the high-dose ANTU effects. Treatment with the low-dose ANTU resulted in an upregulation of gene transcription for keratinocyte growth factor, transforming growth factor-beta, keratinocyte growth factor receptor, and epidermal growth factor receptor as determined through reverse transcription-polymerase chain reaction. According to the authors, these findings suggest that pulmonary cell hyperplasia induced by the low-dose ANTU underlies the resistance to the high-dose ANTU and that the stimulation of hyperplasia may be due to altered growth factor and growth factor receptor expressions (Barton, et al., 2000). The determination of the Clara cell 16-kDa protein (CC16), a small size and readily diffusible protein secreted by the bronchiolar Clara cells, as a biomarker of lung injury in patients with various pathologies (neoplasia, hematologic malignancies, inflammatory pleural effusions, congestive heart failure, hydrothorax, and cirrhosis) as well as in rats exposed to toxicants was evaluated in rats exposed to, among others, ANTU. It was concluded that CC16 present in pleural effusions derived from the serum as well as from the lung following its passage across the visceral pleura (Hantson, et al., 2008; Hermans et al., 1998). (f) Treatment of Poisoning in Animals Mortality of rats caused by 5 mg/kg of ANTU was reduced when allylthiourea, isopropylthiourea, ethylenethiourea, or ethylidenethiourea was administered simultaneously with or a very short time after the ANTU. The first two compounds reduced the survival time of rats that died, but the last two compounds slightly prolonged it (Meyer and Saunders, 1949). Although the reduction in mortality was

Chapter | 100  Rodenticides

statistically significant, the degree of protection was small. Furthermore, these results in the rat may be more closely related to the phenomenon of tolerance than to antidotal action in the usual sense. In any event, Carroll and Noble (1949) found that tolerance to phenylthiourea and ANTU could be produced not only by small dosages of the compounds themselves but also by a number of related and some apparently unrelated compounds. The ability of the effective thiourea-like substances to confer protection was unrelated to their acute toxicities or antithyroid activities. Protected rats failed to develop pulmonary edema or pleural effusion following dosages of toxic thioureas lethal to untreated rats. Thyroidectomized rats could be made tolerant as readily as intact rats. Following a large dose, phenyl­ thiourea was excreted in the urine of a tolerant rat in sufficient quantity to kill a normal animal.

100.4.1.3  Toxicity to Humans (a) Accidental and Intentional Poisoning The absence of a report of uncomplicated poisoning is noteworthy in view of the extensive use of ANTU in Baltimore and some other places and the fact that an occasional bait must have been eaten by children. Several series of cases were reported from France, where chloralose was used either alone for killing crows or rats or in combination with ANTU for killing rats. In one series of 22 cases, all showed some degree of coma and motor agitation, both characteristic of chloralose poisoning; however, more intense pulmonary symptoms were present where ANTU was involved. The low toxicity of ANTU for humans is indicated by the fact that all the patients recovered, although all had ingested the poison with suicidal intent and, therefore, in relatively high dosages (Tempé and Kurtz, 1972). In another series of cases involving chloralose, 14 involved ANTU also, one involved chloralose only, and the presence of or absence of ANTU was not established in the remainder. In addition to the respiratory difficulty that may be present with any coma, 11 of the 14 persons poisoned by a combination of chloralose and ANTU required intubation mainly because of tracheobronchial hypersecretion, and nine of them required artificial respiration. All survived (Favarel-Garrigues and Boget, 1968). The authors characterized the beginning of tracheobronchial hypersecretion as a secretory storm that started early and sometimes suddenly. The secretion was a white froth that, unlike edema fluid, was not sticky or high in protein. The mildness of x-ray changes contrasted with the clinical gravity of the situation. Oxygenation of the blood was always more nearly complete than in acute pulmonary edema. The hypersecretion disappeared rapidly, often in less than 1 h. Apparently, if patients poisoned by a combination of chloralose and ANTU are treated properly, their illness is no more protracted than in poisoning by chloralose alone (see Section 100.1).

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(b) Use Experience Laubstein (1962) reported a case of eczema that he attributed to occupational exposure to ANTU. On the basis that -naphthylamine is an impurity in ANTU, Case (1966b) raised the possibility that persons who distribute ANTU may be in danger of bladder cancer. No epidemiological evidence was offered. Later, Case (1966a) mentioned that an investigation of the occupational history of two rodent operators who were suffering from bladder tumors had revealed that different batches of ANTU differed in the degree of contamination with naphthylamine, and some of the contaminant was -naphthylamine. As a result, the Ministry of Agriculture, Fisheries, and Food recommended in May 1966 that the use of ANTU be restricted to professional operators, and in November 1966 an advisory committee recommended that use of the compound stop until their investigation was complete (Anonymous, 1966). In 1982, Davis et al. reported 14 cases of urothelial tumors observed among 51 rodent operatives exposed to ANTU in the United Kingdom between 1961 and 1980. In the United States as a whole, the age-adjusted death rate for cancer of the bladder increased from 3.1 to 4.1 per 100,000 from 1931 to 1945, when ANTU was discovered, and it continued to increase more slowly until 1953, when it reached 4.4. Since 1953, the values have varied around slightly over 4.3 as a mean. The declining increase in rate from 1945 to 1953 occupied a period less than the average latent period for cancer of the bladder among men with heavy exposure to naphthylamine used in the manufacture of dyes. Thus there is no evidence for any carcinogenic action of ANTU in the general population. Because ANTU had been used so extensively in Baltimore, as described by Richter and his colleagues, the matter was investigated here. Because of the wide fluctuations in rates based on small frequencies, it was not practical to compare death rates for single years; therefore, the data were combined for 3-year periods. For 1949–1965, the rate per 100,000 population varied from the earliest value of 4.4 to 2.9 with no definite trend but certainly with no increase. (c) Dosage Response The threshold limit value is 0.3 mg/m3 of air over an 8-h work shift (OSHA standard). (d) Treatment of Poisoning If treatment were required, it would have to be symptomatic.

100.5  Anti-vitamin K compounds 100.5.1  Overview As reviewed by Link (1944, 1959), knowledge of the antivitamin K compounds began not with vitamin K but with

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OH

O

O

OH

O

O O O Dicoumarol

Coumarin CH3

CH3

O OH

O

O OH

O

O

Warfarin

O O

Coumafuryl O

O

O

O

C CH

C CH O

O Diphacinone O

Chlorophacinone

CH3 O

CH3

Cl

CH3

–

hemorrhagic disease of cattle, which was first recognized in the 1920s on the prairies of North Dakota and neighboring Alberta. It was found that the condition was not caused by a microorganism or a nutritional deficiency but was associated with sweet clover that had gone bad. Hence the condition was known as “sweet clover poisoning.” When cattle or sheep had improperly cured hay made from the common varieties of sweet clover (Melilotus spp.) as their only food, the clotting power of their blood decreased in about 15 days and they often died of internal hemorrhage in 30–50 days. If the disease had not progressed too far, it could be reversed by substituting good hay or by transfusion of blood freshly drawn from normal cattle. Link first learned of the problem in December 1932. During the following February, a Wisconsin farmer came to his laboratory with a dead heifer, a milk can containing blood with no power to clot, 100 pounds of spoiled sweet clover and the all too common, tragic story of cattle dying on an isolated farm. In Link’s laboratory, a practical bioassay for hemorrhagic effect was developed. It was not until June 1939 that the active poison was isolated and crystallized. Using improved methods of isolation developed after the identity of the compound was known, it was shown that the compound was present in spoiled hay at a concentration of about 60 ppm. The structure was shown to be 3,3-methylene-bis(4-hydroxycoumarin), later known as dicoumarol or by the trade name Dicumarol, and it was synthesized in April 1940. The biological synthesis during spoilage of the hay can be rationalized as an oxidation of coumarin (the compound responsible for the characteristic sweet smell and bitter taste of sweet clover) and the subsequent condensation of two molecules of 4-hydroxycoumarin with formaldehyde (see Figure 100.3). When synthetic dicoumarol became available in quantity, the essentials of its pharmacological action were established quickly. Between 1940 and 1942, it was rapidly adopted for treatment of thromboembolic disease in humans. About 50 clinical reports were published between 1941 and 1944. In 1942, Link himself set up field trials to test the suitability of dicoumarol as a rat poison. Tests by O’Connor (1948) using a concentration of 0.44 mg/g were reported as highly successful. However, tests carried out by the U.S. Public Health Service (Hayes and Gaines, 1950) led to the same conclusion as those of Link: that dicoumarol was impractical as a rat poison. While the medical and possible rodenticidal uses of dicoumarol were being explored, over 100 analogues of the compounds were synthesized in Link’s laboratory; they were arranged according to chemical classification and assigned numbers by Overman et al. (1944). In the hope of finding a therapeutic agent other than dicoumarol, the anticoagulant activity of some of those analogues was reappraised using not only rabbits (the species used to detect the hemorrhagic agent of sweet clover poisoning) but also

CH2—CH2–CH2–CH –CH3 3

Vitamin K1 Figure 100.3  Coumarin, dicoumarol, some synthetic rodenticides, and a natural form of vitamin K.

rats, mice, and dogs. Work between 1946 and 1948 identified compounds No. 42 and No. 63 as much more potent than dicoumarol in the rat and dog and as capable of producing a more uniform anticoagulant response and of maintaining a more severe hypoprothrombinemia without visible bleeding than was possible with dicoumarol. Partly on the basis of these observations and partly on the basis of lack of taste and odor, ease of manufacturing the pure compound, and convertibility to a stable water-soluble salt, compound No. 42 was selected. Early in 1948 it was proposed as a rodenticide and promoted by the Wisconsin Alumni Research Foundation. It soon became evident that compound 42 was an important rodenticide. Link (1959) recalled that, although late in 1950 he proposed warfarin for clinical trial, fear of using a highly successful rat poison as a drug prevented significant progress until April 1951, when knowledge of an unsuccessful suicide effectively treated by vitamin K and transfusion of fresh whole blood (Holmes and Love, 1952) brought reassurance. Progress was so rapid that warfarin was used in 1955 for treating then-President Eisenhower.

Chapter | 100  Rodenticides

The use of Dicumarol as a drug and the use of warfarin as a drug and as a rodenticide did not go unnoticed by those who sought a compound even more effective than warfarin – and free of patent restrictions. The result was a number of alternative compounds available either as drugs or rodenticides, or, in the case of diphacinone, used like warfarin for both purposes. The appearance of rats resistant to warfarin and to other early anticoagulant rodenticides has stimulated the search for more potent, fast-acting compounds. These are usually called single-dose rodenticides or second-generation anticoagulants, among which difenacoum and brodifacoum are coumarin derivatives (see Figure 100.4). Anticoagulant rodenticides are divided according to their chemical structure into two main groups: hydroxy­coumarins and indandiones. Indandiones include chlorophacinone, diphacinone, pindone, and valone. Hydroxycoumarin rodenticides are subdivided into first-generation – warfarin, coumachlor, coumafuryl, coumatetralyl – and secondgeneration – brodifacoum, difenacoum, bromadiolone, difethialone and flocoumafen. The latter family of compounds is also known as long-acting anticoagulant rodenticides (LAARs) because of their long half-life in humans: 15–20 days for diphacinone, 120 days for brodifacoum compared to 14–15 hours for warfarin. Coumarin compounds are relatively free of untoward effects when used therapeutically and have been given for long periods without signs of toxicity. LAARs have never been used as therapeutic agents. Anticoagulant rodenticides exert their effects by interfering with the recycling of vitamin K, which is an essential cofactor in the postribosomal carboxylation/activation of clotting factors II, VII, IX, and X by a vitamin K-dependent carboxylase synthesized in the liver. Vitamin K reductase enzymes keep the vitamin in an active (reduced) state. Factors II, VII, IX, and X are proteins serving as enzymatic factors in the common coagulation pathways. They are synthesized in the liver (Petterino and Paolo, 2001; Valchev et al., 2008; World Health Organization, 1995). There are two families of anticoagulant rodenticides: the hydroxycoumarins and the indandiones. The hydroxycoumarins are divided into two categories, the first and the second generation compounds. The indandiones are usually included in the second-generation group because their properties are similar. Both categories of compounds have the same mechanism of action but they differ significantly in their duration of action and response to treatment. The first generation of compounds are considered to be short-acting, requiring multiple exposure to elicit their toxic effects: warfarin for example has a plasma half-life of 14.5 h in the dog and 14.9 h in the mouse compared to a half life of 7–10 days in the rat liver and its anticoagulant effects last for about 1 week. A comparative assessment of the pharmacokinetic properties of intravenously administered sodium warfarin in several species has been performed and reported. The results

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show that its elimination half-life increases in the order: rat monkey  dog  human (Nagashima and Levy, 1969). The second generation of compounds have much longer plasma and liver half-lives: the plasma half-life of diphacinone is 30 days and that of brodifacoum 180 days in rats (Miller, 1984). For bromadiolone, the liver elimination half-life in rat is biphasic with a rapid initial phase of 2–8 days a much slower terminal phase and a complete elimination half-life of 170 days. Similarly, the liver half-life of difenacoum is 118 days (Vandenbrouck et al., 2008). Usually, only a single dose is required to cause bleeding (Mount, 1988). Management of LAARs poisoning has been recently reviewed by Watt et al. (2005) and Caravati et al. (2007). An analytical method has been developed to simultaneously determine and quantify several indandione and hydroxycoumarin active anticoagulant substances in serum and liver: brodifacoum, bromadiolone, chlorophacinone, coumafuryl, coumatetralyl, diphacinone, difenacoum, and warfarin. It is a reversed-phase liquid chromatography with UV detection (Chalermchaikit et al., 1993). A recent analytical method (liquid chromatography– electrospray ionization-coupled mass spectrometry: LC-ESIMS/MS) developed for other applications (determination of veterinary drugs in human and animal food items, for example; Edder et al., 2003; Marek and Koskinen, 2007) was adapted for the simultaneous determination and quantitation of eight anticoagulant rodenticide active substances (brodifacoum, bromadiolone, chlorophacinone, coumatetralyl, difenacoum, difethialone, flocoumafen and warfarin) in plasma and liver tissue (Vandenbroucke et al., 2008). The method fulfills the validation parameters defined by the European Union regulations.

100.5.2  Resistance to Anticoagulant Rodenticides The development of second-generation anticoagulant rodenticides was prompted by the occurrence of rats becoming resistant to the existing compounds. The first observation of such resistance to warfarin in Rattus norvegicus was made in Scotland in 1958 (Boyle, 1960) and later in most of Northern Europe in the late 1960s. Warfarin resistance in the house mouse Mus musculus has followed the same pattern to that of R. norvegicus (Misenheimer et al., 1994). Resistance was also observed in humans treated with coumarin drugs (O’Reilly et al., 1964). The question of the occurrence of resistance in wild rodents in the United States was reviewed as early as 1972 by Jackson and Kaukeinen. Inheritance to warfarin resistance in R. norvegicus is due to the inheritance of an autosomal gene closely related to the gene controlling coat color. The possible mechanisms of underlying coumarin resistance were reviewed and discussed by MacNicoll (1986).

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Considerable amounts of work have been undertaken to elucidate the biochemical and genetic mechanism of rodent resistance to anticoagulant rodenticides. Companies involved in the research and development in this area have created a structure in order to share their knowledge and coordinate their efforts to generate data and information on the subject: the Rodenticide Resistance Action Committee (RRAC). A number of reviews and reports are available dealing with the biochemical and genetic characteristics of the development and maintenance of the resistance to anticoagulants (Buckle et al., 1994; Lasseur et al., 2005; Markussen et al., 2007, 2008; Oldenburg et al., 2007; Pelz et al., 2005; Rost et al., 2004, 2005).

100.5.3  Warfarin 100.5.3.1  Identity, Properties, and Uses (a) Chemical Name 3(-Acetonylbenzyl)-4-hydroxycoumarin is the chemical name. (b) Structure See Figure 100.3. (c) Synonyms The name warfarin (BSI, ICPC, ISO) is in common use except in France, where the compound is called cumafène, in Russia, where it is called zoocoumarin, and in Japan, where it and coumatetralyl both are spoken of as coumarins (JMAF). During development, warfarin was known as Compound-42 or WARF-42. As a drug, the sodium salt is called Coumadin. Trade names for the rodenticide have included Arthrombine-K, Dethmore, and Panwartin. The CAS registry number is 81–81–2. University of Wisconsin biochemistry professor Karl Paul Link and his coworkers first isolated dicoumarin, a molecule in spoiled sweet clover that causes cattle to hemorrhage and die. The discovery led to the synthesis of Dicumarol, the first anticoagulant drug that could be taken orally. The successor to Dicumarol was warfarin. Warfarin is named after WARF, the Wisconsin Alumni Research Foundation, to which Professor Link assigned the patent. (d) Physical and Chemical Properties Warfarin has the empirical formula C19H16O4 and a molecular weight of 308.32. Warfarin is a racemic mixture of R and S stereoisomers. It forms tasteless, odorless, and color­ less crystals with a melting point of 162–164°C. It is practically insoluble in water and benzene, moderately soluble in alcohols, and readily soluble in acetone and dioxane. The sodium salt is fully soluble in water.

Hayes’ Handbook of Pesticide Toxicology

(e) History, Formulations, and Uses The history of warfarin is outlined above. It is formulated as a dust (10 g of active ingredient per kilogram) for use in holes and runs and as a powder (1 and 5 g of active ingredient per kilogram) for mixing with bait to a final concentration of 50 ppm for control of the common rat or 250 ppm for control of the ship rat and mice. Warfarin also is available in many forms of prepared bait.

100.5.3.2  Toxicity to Laboratory Animals Animals intoxicated by warfarin exhibit increasing pallor and weakness reflecting blood loss. Appetite and body weight are not specifically affected. The blood loss may be evident in the form of bloody sputum, bloody or tarry stools, petechiae, or externally visible hematoma. Hematoma formation is more common than free hemorrhage. There is no typical location for hematoma formation, the location of bleeding being apparently a matter of chance in the absence of obvious trauma. Bleeding associated with the central nervous system may be of such location and extent as to cause paralysis of the hindquarters several days before death occurs. Pregnant rats appear slightly more susceptible than nonpregnant ones (Hayes and Gaines, 1950). This may be related to obvious morphological factors, but the decreased metabolism of warfarin in pregnant rats suggests the presence of an inhibitory factor (MacDonald and Kaminsky, 1979). Warfarin may be the only compound for which a log time-log dosage curve with all three segments has been demonstrated experimentally; the 90-day LD50 in rats is only 0.077 mg/kg/day, and the chronicity index is 20.8. Rats tolerated for 300 days a daily dosage slightly greater than the extrapolated 90-day LD 0.01 dosage, specifically 0.02 mg/kg/day. In spite of its considerable cumulative effect, there is a level of intake that is safe for the rat. The same phenomenon permits the use of warfarin as an anticoagulant drug. Other investigators have reported completely different results in the same species. Pyorala (1968), who conducted elaborate studies of the different susceptibility of male and female rats to warfarin, reported LD50 values of 62–102 mg/kg for males and 21–33 mg/kg for females. Hagan and Radomski (1953) reported values of 323 and 58 mg/kg in males and females, respectively. Why these values differ by one or two orders of magnitude from those reported by others is not clear. Warfarin is a racemic mixture whether it is used as a rodenticide or as a drug. Almost all published toxicity figures are for the mixture. However, West et al. (1961) were able to separate the isomers and to determine their absolute configuration. Based on prothrombin time measured 24 h after a single oral dose, the ()(S) isomer was 5.5 times as active as the ()(R) isomer. Based on mortality within 10 days after

Chapter | 100  Rodenticides

starting daily dietary intake, the ()(S)-warfarin was 8.5 times as active as the ()(R) isomer (Elbe et al., 1966). Warfarin is absorbed through the skin of rats, inducing systemic effects. Application of either 0.05 or 0.5 mg/kg bw per day for 3 consecutive days resulted in an increase in prothrombin time. At the high dose, an increase in relative number of granulocytes was also noted. Both doses induced an increased metabolic viability shown by 3-(4,5dimethyl-2-thiazolyl)2,5-diphenyl-2H-tetrazolium bromide reduction. Soluble mediators appear to be involved in the occurrence of those effects since isolated plasma from warfarin-treated rats stimulated nitroblue tetrazolium reduction by granulocytes from nonexposed animals (Kataranovski et al., 2008). These results show that dermal exposure may represent an occupational hazard in manufacturing personnel and bait preparers and applicators. (a) Absorption, Distribution, Metabolism, and Excretion Absorption of warfarin from the skin of rats is slow but measurable. Three dermal doses at the rate of 50 mg/kg had about the same pharmacological effect as three oral doses at 0.6 mg/kg (Sanger and Becker, 1975). Because of either species or formulation differences, the results were very different with guinea pigs and rabbits that received a 0.5% solution of the sodium salt in water (with 8% alcohol and 0.1% of a surface-active agent); single applications at rates of 0.7 and 0.25 mg/kg caused a marked change in prothrombin times in guinea pigs and rabbits, respectively. In fact, one dermal dose at the rate of 0.25 mg/kg was about as effective in rabbits as an oral dose of 2.0 mg/kg (Fristedt and Sterner, 1965). There is great individual variation in the binding of warfarin by the serum proteins of laboratory rats. The rate of excretion showed a strong positive correlation with the concentration of free drug in the plasma (Yacobi and Levy, 1975). Ninety-six hours after intraperitoneal injection of warfarin, the concentrations of activity in the kidney, liver, and pancreas were 3, 12, and 15 times, respectively, greater than that in the blood (Link et al., 1965). The significance of the pancreatic accumulation remains obscure. Warfarin is readily hydroxylated in vitro and in vivo by rat liver microsomal enzymes to form 6-, 8-, and especially 7-hydroxywarfarin. Formation of another metabolite is catalyzed by the soluble fraction of liver in either the presence or absence of oxygen (Ikeda et al., 1968a,b; Ullrich and Staudinger, 1968). Formation of all these metabolites is stimulated by phenobarbital, chlordane, or DDT. The metabolism is a true detoxication. The inducers can increase the LD50 of warfarin by more than 10-fold (Ikeda et al., 1968a). A later study of rats that had received [14C]warfarin revealed the following compounds in the urine: unchanged

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warfarin (6.6%), 4-hydroxywarfarin (21%), 6-hydroxywarfarin (15.4%), 7-hydroxywarfarin (8.9%), a glucuronide of 7-hydroxywarfarin (3.0%), and an intramolecular condensation product, 2,3-dihydro-2-methyl-4-phenyl-5oxo-gamma-pyranol (3.2c)(1)benzopyran (DHG) (6.6%). These metabolites were found in the feces also but in different relative concentrations (Barker et al., 1970). No radioactive carbon dioxide derived from warfarin has been found in exhaled air (Link et al., 1965). Many of the same metabolites were excreted by guinea pigs, but the proportions were different. Salicylic acid, not found in the rat, was found in guinea pig urine. Of all metabolites recovered, only 4-hydroxywarfarin and DHG showed anticoagulant activity. That of 4-hydroxycoumarin was slight. That of DHG showed two peaks, of which the second was stronger. This suggests metabolism of the compound, perhaps back to warfarin (Deckert, 1973). Rats injected intraperitoneally with [14C]warfarin excreted approximately 90% of the activity in 14 days, about half in the urine and half in the feces (Link et al., 1965). Approximately 10% of the activity from [14C]warfarin was excreted in the bile of rats within 5 h after intraperitoneal injection, but little radioactivity appeared in the feces. Nearly all of the metabolites in the bile were conjugated; they could be released with about equal ease by incubation with -glucuronidase of gut flora (Elmer et al., 1977; Powell et al., 1977). The metabolites identified were the same as those found slightly later in the urine. When guinea pigs were injected with 1 or 2 mg of [14C]warfarin, about 50% of the activity was recovered from urine excreted during the first 12 h and 87% was found in urine within 7 days. A smaller percentage of large doses was excreted promptly (Deckert, 1973). The action of warfarin (and of fumarin and coumatetralyl as well) on the smooth muscle of the isolated intestine of the rabbit, of the rat (Rattus rattus and Rattus norvegicus), and of Bandicota bengalensis was studied in vitro. There was a fairly identical reduction in peristaltic activity by all three compounds in all four species. The effect was reversible, thus indicating no permanent damage to the tissue (Renapurkar and Deoras, 1982). Warfarin is bound to albumin but can be displaced from albumin by several compounds, including metals (Brodie, 1964; Chakrabarti, 1978). In humans, the racemic mixture has a half-life of approximately 36–42 h and the S-isomer is five times more potent as a vitamin K antagonist than the R-isomer (Breckenridge et al., 1974). Absorption of warfarin is rapid and complete. It is highly protein bound (98%), primarily to albumin; only the free form is pharmacologically active (O’Reilly, 1969). The hepatic metabolism of the two isomers has clinically significant implications for drug interactions. The S-isomer is primarily metabolized by cytochrome P450 2C9 and to a lesser degree P450 3A4 and

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is eliminated in the bile. The R-isomer is primarily metabol­ ized by cytochrome P450 1A2 and 3A4 and is excreted in the urine as inactive compounds (Jaffer and Bragg, 2003). (b) Resistance to Warfarin Genetic resistance to warfarin among rodents, lagomorphs, and humans is discussed in Section 100.3.1.3. Two cases of intriguing warfarin resistance in humans were reported by Kempin (1983). Both patients under anticoagulant therapy could not be kept within therapeutic range. The common factor that was found was heavy daily intake of broccoli (250–450 mg/day). Broccoli is an important dietary source of vitamin K (200 g/100 g). When the vegetable was removed from the diet, the anticoagulant therapy became effective. It is interesting to note that a patent has been granted for an “invention” whose objective is to provide a rodenticidal composition enhancing the toxicity of warfarin in rodents (WO/1999/045778). This invention is based on the hypothesis according to which “resistance” to warfarin in rodents may not be true resistance but rather the result of bacterial action that rapidly metabolizes warfarin in the gastrointestinal tract of rodents. Therefore, adding an antibiotic in the bait formulation reduces the degradation of warfarin and consequently enhances its efficacy. Any antibiotic that is effective against the warfarindegrading bacteria is suitable. These include tricyclic and tetracyclic antibiotics such as tetracycline and its salts and derivatives. Several other chemical families may also be used. Only a relatively small amount of antibiotic is required, generally ranging from 0.05 to 1.0% based on the bait weight. This observation is to be related to the drug interactions that have been reported in humans (see below). The hypothesis according to which anticoagulant resistance to warfarin could be resulting from a rapid degradation by gastrointestinal bacteria appears to be untrue since it was later demonstrated that this resistance was genetically determined (see below), at least in humans and in rodents. Berny et al. (2006) have studied the rate and extent of ruminal degradation of warfarin (as well as chlorophacinone and bromadiolone) in vitro and evaluated their oral bioavailability as well as their clinical and hemostatic effects in adult sheep. Samples of ruminal fluid were incubated with a given quantity of each compound in order to determine the kinetics of the ruminal degradation over a 24-h period. The animals were fitted with a rumen cannula. To establish the plasma kinetics of the anticoagulants, each animal received the test material either by IV injection or via the cannula at 2-month intervals. Each animal received three administrations. At 10- to 15-day post-treatment intervals, prothrombin time (PT), plasma concentration, and clinical signs were assessed. In the rumen extracts in vitro, anticoagulants were only slightly degraded: less than 15% over 24 h. In vivo, PT was prolonged (80 s), but no clinical

Hayes’ Handbook of Pesticide Toxicology

signs of coagulopathy were observed. PT returned to normal within 2 weeks. In the animals, the bioavailability of all three compounds was high, 79% for warfarin, 92% for chlorophacinone, and 88% for bromadiolone. According to the authors, the kinetics of these compounds in sheep and in other mammals is similar. They concluded that the lack of sensitivity of ruminants to these anticoagulants could not be attributed to a ruminal degradation process or particular toxicokinetic properties. (c) Biochemical Effects Warfarin has two actions: inhibition of synthesis of vitamin K-dependent factors (VII, proconvertin; IX, Christmas factor; and X, Stuart factor) and decrease of the production of prothrombin (factor II) in the liver (Coon and Willis, 1972). In addition, warfarin induces capillary damage. There is unconfirmed evidence that these two actions are produced by the two moieties of the molecule. Thus 4-hydroxycoumarin inhibits the formation of prothrombin and reduces the clotting power of the blood, whereas there is some evidence that at sufficient dosage benzalacetone produces capillary damage and leads to bleeding upon the very slightest trauma. Significantly enough, vitamin K has an antidotal action against both actions of warfarin up to a certain point (Varon and Cole, 1966). The basis for the change in vitamin metabolism associated with poisoning and the alteration of this metabolism in resistant animals probably involves a warfarin-binding protein in the microsomal membranes of the liver. Thierry et al. (1970) found that ribosomes isolated from the livers of resistant rats bind only one-third to one-fifth as much warfarin as ribosomes from normal rats, regardless of whether warfarin is injected before the rats are killed for study or is added to the in vitro preparation. Lorusso and Suttie (1972) found that, when [14C]warfarin at a concentration of 0.786 ppm was incubated with microsomal preparations, the concentrations reached were 42.0 and 17.7 pmol/mg of protein, depending on whether the preparations were prepared from normal or from warfarin-resistant rats, respectively. Furthermore, the warfarin was bound firmly to membranes of normal rats but loosely to those of warfarin-resistant rats. Vitamin K deficiency caused a 24% increase in the amount of warfarin bound, but this was overcome in animals given vitamin K 1 h before being killed for in vitro study. Warfarin binding in vitro was reduced 90% in animals injected with warfarin 22 h before being killed. Although the binding protein was a part of microsomal membranes, it seemed unrelated to cytochrome P450. A protein which may be the same as the one just discussed has been isolated and shown to have a molecular weight of about 30,000. It may become adherent to ribosomes in the course of their preparation for biochemical study (Searcey and Graves, 1976). Binding of warfarin to cytochromes P450 and P448 occurs also and may help to explain changes in the rate of

Chapter | 100  Rodenticides

metabolism of warfarin following induction of microsomal enzymes by phenobarbital and other compounds. The stereo-­ chemical aspects of the metabolism of warfarin have been studied in great detail (Kaminsky et al., 1976; Pohl et al., 1976a,b, 1977). Formation of 7- and 8-hydroxywarfarin is promoted by other cytochromes. The same type of cytochrome is mainly responsible for the formation of each corresponding metabolite, regardless of how the activity of liver microsomal enzymes has been induced (Fasco et al., 1979). Warfarin has been reported to inhibit (Biezunski, 1970) or to promote (Beracki and Bosmann, 1970) the synthesis of liver microsomal protein and other liver protein. The contradictory results may be explained by differences in procedure, but exactly how is unclear. It is also unclear what bearing the results have on the pharmacological action of warfarin. Warfarin causes a relative increase in vitamin oxide in the plasma or liver of people (Shearer et al., 1973) and rats (Matschiner et al., 1970). The oxide is a naturally occurring compound. In vitamin K-deficient but otherwise normal rats, the oxide and vitamin are equally effective, but the oxide is not therapeutic in warfarin-treated rats. It has been proposed that coumarin and related anticoagulants act by inhibiting the conversion of the oxide back to the active vitamin and that the oxide per se is inhibitory. Involvement of the vitamin–vitamin oxide cycle in the action of warfarin seems very likely, since the effect of warfarin on this cycle is greatly reduced in resistant rats. The hypothesis that warfarin inhibits prothrombin synthesis by causing accumulation of the oxide does not appear tenable (Caldwell et al., 1974). However, it seems likely that the brevity of the action of vitamin K in the treatment of poisoning is the result of its irreversible conversion to the epoxide (Shearer and Barkhan, 1979). The superiority of vitamin K over vitamin K3 in treating warfarin poisoning has been established experimentally (Penumarthy and Oehme, 1978). For more detail than can be discussed here regarding vitamin K and vitamin K-dependent proteins a book edited by Suttie (1979) is available. Although l-histidine at a dietary level of 400 ppm was without effect on rats, it potentiated the lethal action of warfarin (50 ppm) in both laboratory and field tests (Rao, 1979). The biochemical basis for this action of histidine should be explored. (d) Effects on Organs and Tissues A possible oncogenic effect of prolonged warfarin therapy was speculated by Krauss (1982) based on a report by Gore and associates (1982) of an increased incidence of cancer in patients occurring 3 or more years after a pulmonary embolism. This highly speculative deduction made from a very limited number of cases has been criticized by Zacharski (1982) who cited results from two cohort studies

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(Annegers and Zacharski, 1980; Michaels, 1974). In those studies, no increased incidence of malignancies was observed in patients who had received long-term anticoagulant treatments and had been followed for several thousand patient-years. Warfarin has been established as a true teratogen in several ����������������������������������������������������� animal ���������������������������������������������� species, inducing several types of anomalies: embryotoxicity (hemorrhagic fetal syndrome, metabolic fetal liver damage, growth retardation), hind limb malformations, tibial growth plate abnormalities, maxillofacial hypoplasia, hydrocephalus, etc. (Feteih et al., 1990; Howe and Webster, 1992; Kronick et al., 1974). (e) Pathology Animals killed by warfarin show most extreme pallor of the skin, muscles, and all the viscera. In addition, evidence of hemorrhage may be found in any part of the body but usually only in one location in a single autopsy. Such blood as remains in the heart and vessels is grossly thin and forms a poor clot or no clot. In a report of a boxer dog poisoned with a mixture of warfarin and calciferol, delayed necrosis of the tip of tongue and large areas of necrosed skin were seen. It is difficult, however, to attribute the damaging effect on the walls of the vessels to warfarin alone, since calciferol also can induce such lesions (Edlin, 1982). (f) Treatment of Poisoning in Animals A diet containing selenium at a concentration of 2.5 mg/kg of feed was protective against the toxic effects of aflatoxin B1 (a bifuranocoumarin) and warfarin in pigs given in four daily oral doses of 0.2 mg/kg of body weight. Selenium is a component of glutathione peroxidase, an enzyme that prevents the production of free radicals (Davila et al., 1983). The treatment of choice of anticoagulant rodenticide poisoning is vitamin K1. Depending upon the circumstances, treatment may be administered either orally or subcutaneously (never intravenously because occasional anaphylactic reactions have been reported). When given orally, vitamin K1 should be given with food, preferably fat-containing, in order to enhance its absorption. However, it is recommended to administer the first dose subcutaneously (0.25–2.5 mg/kg bw), followed by oral administration twice a day at half the initial dose for the following 7 days. Coagulation parameters should be monitored weekly until values remain stable for 5–6 consecutive days (The Merck Veterinary Manual, 2008).

100.5.3.3  Toxicity to Humans (a) Experimental Exposure When nine normal men and five normal women were given a single oral dose of warfarin at the rate of 1.5 mg/kg, maximal concentration in plasma was reached in 2–12 h. Maximal depression of prothrombin activity was between 36 and 72 h. Their individual increases in prothrombin time

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were proportional to their half-time for disappearance of warfarin from the plasma. In other words, the pharmacological effect was greatest in those with slower excretion. The half-times for disappearance from the plasma varied from 15 to 58 h with a mean of 42 h. Absorption of warfarin from the gastrointestinal tract was apparently complete; no warfarin was found in the stool even after massive doses, and plasma levels and prothrombin activity responses were virtually identical following oral and intravenous administration at the same rates (O’Reilly et al., 1963). Having established the absolute configuration of the four warfarin alcohols, Chan et al. (1972) administered them to volunteers. Reduction of the alcohols was stereoselective. The rate of elimination of one of the isomers (R,S) was much slower than that of the others, and its effect was more sustained. The resulting metabolites were biologically active but not as active as warfarin itself. Six normal subjects were given a single dose of warfarin at the rate of 1.5 mg/kg. Three weeks later, the same people were given 200 mg of phenylbutazone three times a day for at least 8 days; on the 4th day, warfarin was repeated at 1.5 mg/kg. Compared to warfarin alone, administration of warfarin with phenylbutazone increased the prothrombin time even though the plasma concentration and biological half-life decreased. The result (in the face of an obvious inactivation of warfarin) was attributed to displacement of warfarin by phenylbutazone from binding to plasma albumin, making more free drug momentarily available to receptor sites in the liver (O’Reilly and Aggeler, 1968). The mutual displacement of phenylbutazone and warfarin from human plasma albumin has been studied in vitro (Solomon et al., 1968). As shown by a study in seven volunteers, the action of triclofos (trichloroethyl sodium phosphate) is similar to that of phenylbutazone. A dosage of triclofos at the rate of 22 mg/kg/day prolonged the prothrombin time even though the dosage of warfarin was reduced. Trichloroacetic acid, a metabolite of triclofos, accumulated in the plasma to an average concentration of 80 ppm. The displacement of warfarin from albumin by trichloroacetic acid was sufficient to account for the observed potentiation of warfarin (Sellers et al., 1972). At least in the rat, sodium salicylate has a similar effect (Coldwell et al., 1974), but phenobarbital does not significantly influence the binding capacity of the plasma for warfarin (Ikeda et al., 1968a). In a similar study with 10 male volunteers, both phenobarbital and glutethimide lowered the plasma warfarin concentration and reduced the half-life of warfarin by nearly 50%; chloral betaine had a slight effect also. Phenobarbital and glutethimide significantly reduced the hypoprothrombinemia response of warfarin, but results with chloral betaine were indistinguishable in this regard from results for placebotreated and untreated controls (MacDonald et al., 1969). The effects of cimetidine (a drug used to treat peptic ulcer) on the kinetics and dynamics of warfarin were studied in seven volunteers. It was shown clearly that

Hayes’ Handbook of Pesticide Toxicology

cimetidine acted by inhibition of drug metabolism without significant effect on the binding of warfarin to plasma protein (Breckenridge et al., 1979). (b) Therapeutic Use Therapeutic use of anticoagulant drugs in humans has been thoroughly reviewed and standardized on the basis of their mechanism of action and the normalized index INR (International Normalized Ratio) that was developed in 1982 to take into account the variations of thromboplastin sensitivity and different ways of reporting the prothrombin time across the world (Hirsh et al., 2001). Use of warfarin as a drug offers greater dosage and, therefore, greater opportunity for side effects than pest control operators encounter. Of course, bleeding is the most common complication of treatment. Most of it is clinically insignificant. Probably many cases remain unpublished. According to one study, the incidence among hospitalized patients was 10%, and it was 40% among ambulatory patients. The incidence of serious hemorrhage was estimated at 2 and 10% in hospitalized and ambulatory patients, respectively. A series of case reports illustrated some of the circumstances leading to serious hemorrhage. It was concluded that the complication can be kept at a minimum by careful selection of patients, informed and adequate supervision by the physician, and reliable laboratory control (Pastor et al., 1962). Although the diagnosis in most cases of hemorrhage is obvious, there are exceptions. For example, two cases of intestinal hemorrhage leading to an initial diagnosis of acute abdomen have been reported (Cocks, 1960). Macular, papular, pruritic, or vesicular rashes due to warfarin are unusual, but those that do occur often are in patients who had taken the drug without untoward effect for 3 or more months. The skin returned to normal slowly after medication was stopped, but the dermatitis recurred within 2 or 3 days when medication was renewed (Schiff and Kern, 1968). Necrosis of the skin and subcutaneous tissues of localized areas has been attributed to warfarin only rarely and not always convincingly. For example, in a case reported by Vaughan et al. (1969), a totally unexplained illness suggestive of but not proved to be thrombophlebitis and pulmonary embolism preceded warfarin therapy by 6 weeks and may have been the underlying cause of the complication attributed circumstantially to warfarin. In two other cases, a more persuasive interpretation was made, namely that anticoagulants (heparin and warfarin) acted as neither the preparatory nor the provoking substance for the localized necrosis attributed to their use, but that underlying disease processes, including intravascular coagulation, sepsis, or localized inflammation, triggered a localized purpuric reaction that was then intensified by the warfarin therapy (Martin et al., 1970). An entirely different kind of complication has involved potentiation of warfarin by disulfiram (Rothstein, 1968)

Chapter | 100  Rodenticides

or interference with its action by other drugs including griseofulvin (Cullen and Catalano, 1967) and phenobarbital (Robinson and MacDonald, 1966) or by insecticides. In one case, medical use of warfarin was nullified by use of 5% toxaphene and 1% lindane to dust sheep. Response to warfarin returned to normal within about 3 months after exposure to the insecticides (Jeffery et al., 1976). Of course, discontinuing a drug that promotes the metabol­ ism of warfarin has the same effect as introducing a drug such as disulfiram that interferes with metabolism of the anticoagulant. The success of cardiac valve prostheses requires anticoagulation to prevent immobilization of the valve by thrombi and to minimize the chance of emboli. Installation of such prostheses in young women to correct rheumatic mitral or aortic stenosis increases their chance of surviving and reproducing, but it necessarily complicates any pregnancies they may have. Anticoagulants may also be used in women of childbearing age to treat thrombophlebitis, embolic disease, and a few other conditions. It has long been recognized that administration of any anticoagulant during pregnancy increases the danger of hemorrhage either during the course of gestation or during delivery. It gradually has become evident that warfarin also is teratogenic in humans (Beckert et al., 1975; DiSaia, 1966; Keber et al., 1968; Pettifor and Benson, 1975; Shaul et al., 1975; Sherrod and Harrod, 1978; Tejani, 1973). The first described case of multiple congenital anomalies in an infant associated with maternal ingestion of warfarin during pregnancy was reported by DiSaia in 1966. At least 29 cases of congenital anomaly have been attributed to warfarin (Hall et al., 1980). Most if not all of the recognized cases have involved nasal hypoplasia ranging from barely recognizable to very severe. Many of the babies had chondrodysplasia punctata, and this defect of cartilage development may be the basis not only of the nasal deformity but also of defects of the bones such as meningocele, deformities of the limbs, and a high arched palate seen much more rarely in babies of women treated with warfarin during the first trimester. Other teratogenic effects reported in one or more cases include microphthalmia, blindness, hydrocephalus, persistent truncus arteriosus, and mental retardation. Of 423 reported pregnancies in which coumarin derivatives were used, over two-thirds resulted in apparently normal infants, one-sixth resulted in abortion or stillbirth, and one-sixth resulted in abnormal liveborn infants of which 29 showed fetal embryopathy. The critical period of exposure seemed to be between 6 and 9 weeks of gestation. Five cases of typical embryopathy and eight other cases showed central nervous system abnormalities following exposure to coumarin derivatives during gestation, but no critical period of exposure was evident (Hall et al., 1980; Stevenson et al., 1980).

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Twenty-nine of 423 pregnancies is a high incidence of teratogenic effect even if one takes into account that only medically reported cases were available for consideration. A different kind of evidence for the teratogenic action of warfarin involved a family with no history of consanguinity, birth defects, or mental retardation that produced one normal child in a pregnancy without warfarin and two deformed children in separate pregnancies in which warfarin was used (Sherrod and Harrod, 1978). On the other hand, inasmuch as defects occur in only about one-third of instances, none may be found in some small series of cases (Chong et al., 1984). The use of heparin during gestation does not result in a significantly better outcome of pregnancy than that obtained with warfarin. In 135 published cases, about twothirds were apparently normal, one-eighth were stillborn, and one-fifth (of whom one-third died) were premature (Hall et al., 1980). Kaplan (1985) and Zakzouk (1986) reviewed the subject of warfarin-associated malformations and established that, based on the timing of warfarin exposure, secondand third-trimester exposure predisposes to central nervous system abnormalities, whereas first-trimester exposure is associated with the warfarin embryopathy: midface and nasal hypoplasia, optic atrophy, hypoplasia of the digits, and mental impairment. In addition, Kaplan reports a case of Dandy-Walker malformation associated with warfarin exposure confined to weeks 8–12 of gestation. Four previous cases of such association were known but the gestational exposure period was much longer. Two recent surveys and reviews have established direct links between early in utero exposure to anticoagulant compounds and embryotoxicity malformations (Schaefer et al., 2006; van Driel et al., 2002). Not only has hereditary resistance of people to warfarin been observed (O’Reilly, 1970), but also exceptional susceptibility, also presumably on a hereditary basis, has been reported (Solomon, 1968). On the other hand, the case of a patient on warfarin treatment with a bleeding complication while his prothrombin time was in the normal therapeutic range was reported (Chu et al., 1996) with the partial thromboplastin time appearing disproportionally prolonged. On specific factor assay, the factor IX clotting activity was less than 1% of normal. This suggested that his factor IX had increased sensitivity to warfarin. A mutation predicting a substitution of an Ala to Thr (at the 10 position in the propeptide region) was found. This observation has been commented upon by Thompson (1996), who concluded that this mutation represented a variant that predisposed the patient to warfarin toxicity. Several reports of drug interactions with warfarin have been published. Those interactions are quite numerous and varied. For example, the case of a patient with clarithromycin– warfarin interaction (in addition to a clarithromycin–digoxin

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interaction) is reported by Gooderham et al. (1999). Clarithromycin is essentially used to eradicate Helicobacter pylori responsible for gastric ulcer. Cases of interactions with a very large spectrum of drugs and natural substances have been published. These potential interactions must be taken into consideration in patients under warfarin therapy (Feldstein et al., 2006; Holbrook, et al., 2005; Jaffer and Bragg, 2003; Riechelmann and Saad, 2006; Wells et al., 1994). Interactions between warfarin and influenza vaccination have been reported in elderly patients (Poli et al., 2002). The vast majority of those warfarin interactions are related, through either their induction or inhibition, to various forms of isoforms of cytochromes P450, essentially CYP3A4 and CYP2D6, but also CYP1A2, 2B6, 2C19, and 2E1, depending upon the therapeutic classes of compounds (Anonymous, 2002). (c) Accidental and Intentional Poisonings A 32-year-old man was murdered by feeding him warfarin for 13 days. On the 4th day after intake started, the victim began having severe nosebleeds. Later, he bled from the mouth. Two days before death, he complained of pain in his limbs. His symptoms became worse and he died of circulatory failure on day 15 (Pribilla, 1996). The initial symptoms in an attempted suicide using warfarin were back pain and abdominal pain. The onset occurred 1 day after the sixth daily dose. A day after onset, vomiting and attacks of nose bleeding occurred. On the 2nd day of illness, when admitted to hospital, the patient was observed to have a generalized petechial rash (Holmes and Love, 1952). In Korea, a family of 14 persons lived for a period of 15 days on a diet consisting almost entirely of corn (maize) meal containing warfarin. The first symptoms appeared 7–10 days after the eating of warfarin was begun. Massive bruises or hematomata developed at the knee and elbow joints and on the buttocks in all cases. Extensive gum and nasal hemorrhage usually appeared about a day later, and by day 15 blood loss was extensive (Lange and Terveer, 1954). A suicidal gesture that was reported and treated after only a single ingestion of a heaped tablespoonful of 0.5% warfarin produced no illness and not even an increase in prothrombin time (Kellum, 1952). There have been at least two attempted murders with warfarin (Ikkala et al., 1964; Nilsson, 1957). In each instance, there were recurrent bouts of hemorrhagic difficulty, including hematuria, epistaxis, severe bruises without any history of trauma, and intestinal hemorrhage. Abdominal or back pain was present. Each patient recovered promptly in hospital, but one had relapses for nearly a year and the other had bouts of poisoning of 2.5 years following repeated doses. An anticoagulant drug was suspected from the outset in both cases, but finding the source proved difficult. The solution to each case was essentially

Hayes’ Handbook of Pesticide Toxicology

epidemiological, but measurement of warfarin in the plasma was decisive in one case. Faced with the evidence, the daughter-in-law of a 72-year-old woman and the wife of a 69-year-old man both confessed to the police. Although numerous accidental ingestions by children and adults have been reported to the New York Poison Control Center, no known injury from these ingestions has been observed (Jacobziner and Raybin, 1960). An outbreak of hemorrhagic disease due to the use of warfarin-contaminated talcum was described in Vietnam (Martin-Bouyer et al., 1983). Of the 741 cases located in Ho-Chi-Minh City, all in infants (55% under 2 months of age), 177 died. Eleven samples of baby powder were analyzed and concentrations of warfarin ranged from 1.7 to 6.5%. The percutaneous penetration of warfarin contained in the contaminated talc was studied in a young healthy female baboon, treated twice daily with a topical application of 3 g of talc containing 3% warfarin (188 mg/kg/day of warfarin). One control animal was treated with uncontaminated talc. On the fifth day, the treated animal began to show signs of intoxication with profuse bleeding and died. At necropsy there were two large subcutaneous hematomata on the skull and the peritoneal cavity was filled with unclotted blood. On day 3 of treatment, a blood sample showed severe disturbances of the hepatic coagulation factors. Electron microscopy showed an increased number of swollen and misshapen mitochondria in the hepatocytes (Dreyfus et al., 1983). Warfarin was administered to an 11-month-old baby girl by her psychologically disturbed mother. Upon hospitalization the parameters of coagulation were elevated (prothrombin time 53s, control 12s); the child had multiple hematomata and a bloody discharge from her left ear. Treatment with vitamin and infusion of fresh frozen plasma stopped the bleeding (White, 1985). (d) Use Experience The safety record of warfarin used as a rodenticide has been excellent. One case of poisoning has been attributed to extensive, prolonged skin contact in the process of preparing and distributing baits. Unlike most solid bait, which is prepared by mixing ground grain with starch containing warfarin powder, this bait was prepared by pouring a 0.5% solution of the sodium salt over dried bread. The hands of the 23-year-old farmer who used this method were wet with the solution each of the 10 times he made bait during a 24-day period, and he did not wash his hands until several hours after each application. Two days after the last contact with rodenticide, gross hematuria appeared. The next day, hematomata were noticed on the arms and legs; there was dull pain in both groins. The hematuria subsided after 3 days of rest but recurred along with nose bleeding when the man returned to work. When he was admitted to hospital, prothrombin, clotting, and bleeding times were abnormally long and anemia was

Chapter | 100  Rodenticides

severe (hemoglobin, 8.2%; red cell count, 2.9 million/mm3). The patient responded promptly to treatment with vitamin K1 (Fristedt and Sterner, 1965). (e) Atypical Cases of Various Origins Mogilner et al. (1974) described three fatal cases of what they considered to represent Reye’s syndrome. One of the young children was said to have ingested warfarin that she found by the road on her way to kindergarten 4 days before hospital admission; warfarin was found in the urine and traces were present in the blood. In another case, “significant amounts of warfarin were found in the urine and also in autopsy samples of liver and kidney tissue;” no source of the rodenticide was reported, and no information was given that would permit evaluation of the validity of the chemical analysis. There was no indication of warfarin in the third case. The authors considered that it was justified “to add warfarin to the long and varied list of etiological or precipitating factors in Reye’s syndrome.” The death of a 2-year-old boy was attributed to warfarin poisoning on his death certificate but probably was the result of beating, with pneumonia as a terminal event (Hayes and Vaughn, 1977). A patient developed a severe case of exacerbation of his thyrotoxicosis and warfarin toxicity shortly after radioactive iodine therapy for his Graves disease. The levels of thyroid hormones have an influence on the warfarin requirements as hyperthyroid patients are more sensitive to warfarin and thus require lower dosages of warfarin than when they are euthyroid. Close monitoring of these patients is necessary to prevent severe and potentially lethal occurrence of thromboembolic or hemorrhagic side effects (Westphal, 2008). (f) Dosage Response A total dose of 1000 mg of warfarin consumed in 13 days (about 1.1 mg/kg/day) was fatal (Pribilla, 1996). Serious illness followed the ingestion of 1.7 mg warfarin/kg/day for 6 consecutive days with suicidal intent. This would correspond to eating almost 1 pound of bait (0.025% warfarin) each day for 6 days. All signs and symptoms were caused by hemorrhage and, following multiple small transfusions and massive doses of vitamin K, recovery was complete (Holmes and Love, 1952). In the Korean cases, the dosage of the different individuals was determined to vary from about 1 to 2 mg warfarin/kg/day. As a result of this exposure and without benefit of treatment, 2 of the 14 persons died. A 19-year-old girl, who was in a state of shock and severe hemorrhage 2 days after the warfarin diet was discontinued, recovered following a blood transfusion and small daily doses of vitamin K. The remaining 11 members of the family recovered within a week after exposure, although only small daily doses of vitamin K were given and they all had shown marked signs of poisoning when they first accepted treatment. There was reason to think that those who died had

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received slightly higher dosages than those who survived (see Table 2.9 in Hayes, 1975). Recovery of the 12 survivors was complete. The entire episode was made possible only by a series of unusual events and by the extraordinary apathy of the family, resulting in their totally ignoring unmistakable signs of illness (Lange and Terveer, 1954). A single intravenous therapeutic dose of the sodium derivative (40–60 mg or about 0.7 mg/kg) in humans may produce some increase in prothrombin time within 2 h and usually produces a substantial increase within 14 h. The average maximal response is on the 4th day. Spontaneous recovery to normal occurs about 8 days after a single therapeutic dose. Thus significant depression of prothrombin level is maintained for 3–6 days. In the treatment of thromboembolic disease, a maintenance dose of about 2–10 mg/day is required to keep the prothrombin level between 10 and 30% of normal. Patients have been thus maintained for years. If human susceptibility to warfarin were different (as it is in a few genetically determined cases), the therapeutic dosage could be adjusted accordingly. It is interesting, however, that the upper limit of the usual maintenance dosage for humans (about 0.14 mg/kg/day) is an LD95 for the rat. The inherently lesser susceptibility of humans to the compound undoubtedly contributes to its safety as a rodenticide. The threshold limit value of 0.1 mg/m3 indicates that occupational intake of warfarin at the rate of 0.014 mg/kg/day is considered safe. (g) Laboratory Findings Metabolites of warfarin including 5-, 6-, 7-, and 8-hydroxywarfarin and two aliphatic side-chain alcohols have been identified in the urine of normal volunteers who had received a single oral dose at the rate of 1.5 mg/kg (Chan et al., 1972; Lewis and Traeger, 1970). This is, however, not a common way to confirm poisoning. Adequacy of treatment with warfarin usually is followed by measuring prothrombin time. In cases of poisoning, the prothrombin time is greatly prolonged. The coagulation time is definitely increased by the Lee-White method and slightly increased by the capillary tube method. Bleeding time often is normal. Urine may be normal in appearance but may contain many red cells on microscopic examination, or it may be grossly hemorrhagic. The red cell count and hemoglobin gradually fall if bleeding continues. In terminal cases a state of shock develops. Sixty-nine euthyroid patients being treated with warfarin for thromboembolic disease showed no evidence of hyperthyroid condition, and 14 of them showed a hypothyroid tendency associated with an elevation of the thyroxine-binding capacity of plasma globulin. However, no clinical evidence of thyroid dysfunction was reported, and it was uncertain whether the small changes in laboratory tests of thyroid function were caused by warfarin (Braverman and Foster, 1969). Plasma levels of warfarin were 6.8 and 11.2 ppm 4 and 7 h, respectively, after the ingestion of 50 mg of warfarin sodium in a suicide attempt. Plasma levels declined

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thereafter, and the half-time disappearance was calculated as 46 h. Part of the dose was removed by gastric lavage soon after ingestion. This and other appropriate treatment prevented any increase in bleeding tendency (Cole and Bachmann, 1976). (h) Pathology Apparently there is only one complete description of human pathology associated with uncomplicated warfarin poisoning, that of Pribilla (1996). The findings in that case were strikingly different from typical findings in the rat: exsanguination was less complete, as indicated by the fact that the liver was not tan in color and bleeding was far more generalized and not restricted to one or a few large hematomata. The two factors may be related in that bleeding into the organs may have interfered with their function and hastened death. In addition to generalized bleeding (due mainly to deficiency of coagulation), evidence of capillary damage and of parenchymal injury of the liver was found in the human case. In spite of the obvious differences between the findings in human and rat, the similarity was also striking because subserosal and intraseptal bleeding was prominent in the human case. (i) Treatment of Poisoning After blood has been taken for prothrombin and other differential diagnostic tests, vitamin K in a dose of 5–10 mg should be given three times on the first days of treatment irrespective of symptoms. The vitamin should be given intravenously slowly, usually by infusion. Smaller doses should be continued until the prothrombin time has reached normal. In a seriously ill patient, a small transfusion of carefully matched whole blood should be given initially and repeated daily until the patient has returned to normal. Such a patient should also be given vitamin K. If it were ever necessary to treat a patient in shock from blood loss resulting from warfarin poisoning, frequent small transfusions and a complete consideration of the blood chemistry would be in order. Any large hematomata should be the subject of a surgical consultation, but any surgical action should be taken only after the clotting power of the blood is restored to normal. The progress of the patient should be followed by the prothrombin test. Tests should be made at least twice daily until a return to normal is clearly established and stable.

100.5.4  Coumafuryl 100.5.4.1  Identity, Properties, and Uses (a) Chemical Name 3-[-1(2-Furanyl)-3-oxobutyl]-4-hydroxy-2H-1-benzopyran-2-one is the chemical name. (b) Structure See Figure 100.3.

(c) Synonyms Coumafuryl is the common name approved by ISO. Other names for the compound include fumarin (BSI) and tomarin (Turkey). Trade names include Fumarin, Fumasol, Krumkil, Lurat, Ratafin, Rat-a-way, and Tomarin. The CAS registry number is 117-52-2. (d) Physical and Chemical Properties The empirical formula for coumafuryl is C17H14O5, and the molecular weight is 298.28. It is a crystalline solid melting at 214°C. (e) Use Coumafuryl is an anticoagulant rat poison.

100.5.4.2  Toxicity in Laboratory Animals Coumafuryl is very similar to warfarin (see Figure 100.3). The oral LD50 for rats is quoted as 0.4 mg/kg (Wiswesser, 1976). Coumafuryl poisoning was observed in young chicks (less than 1 week old) with a mortality rate of 100%. Hemorrhage and unclotted blood were noted in the abdominal and thoracic cavities. At necropsy, crops and gizzards contained feed. Analysis of the content detected approximately 340 mg/kg of coumafuryl. Investigations found coumafuryl was present in the wood-straw mats in the chicken boxes (Munger et al., 1993).

100.5.4.3  Toxity to Humans It is inevitable that a number of children and perhaps others have ingested coumafuryl. In fact, McLeod (1970) reported that coumafuryl and warfarin were among the pesticides most often ingested by persons (mainly children) admitted to a large hospital in New Orleans. However, these two compounds were the least hazardous pesticides in terms of morbidity. (a) Treatment of Poisoning Treatment is the same as that for warfarin (see Section 100.5.3.3).

100.5.5  Diphacinone 100.5.5.1  Identity, Properties, and Uses (a) Chemical Name 2(Diphenylacetyl)indan-1,3-dione is the chemical name. (b) Structure See Figure 100.3. (c) Synonyms Diphacinone (ANSI, BSI, ISO) is the common name in use except in Turkey and Italy, where diphacin is used, and in

Chapter | 100  Rodenticides

Russia where ratindan is used. Other nonproprietary names include dipazin and diphenacin. As a drug, the compound is known as diphenadione. Trade names for formulated baits containing diphacinone include Diphacine, Ramik, Promar, and Gold Crest. The CAS registry number is 82-66-6. (d) Physical and Chemical Properties Diphacinone has the empirical formula C23H16O3 and a molecular weight of 340.40. Its technical grade is a yellow crystalline powder melting at 145°C; it is slightly soluble in water (0.3 mg/l) and soluble in acetone (29 g/l) and toluene (73 g/l). Diphacinone is rapidly decomposed in water by sunlight. (e) History, Formulation, and Uses The rodenticidal activity of diphacinone was described in 1952. It is formulated as prepared weather-resistant baits (pellets or meal) in concentrations of 50 mg/kg. A dry concentrate (1 g/kg) for mixing with cereal bait is also available. All formulations are for professional application only. Diphacinone is used to control mice, rats, prairie dogs (Cynomys spp.), ground squirrels, voles, and other rodents.

100.5.5.2  Toxicity in Laboratory Animals In a study of bishydroxycoumarin, ethyl biscoumacetate, and 17 analogues of indandione in rabbits, diphacinone was found to be the most hypoprothrombinemic. A marked response lasting about 7 days was produced by a dosage of only 0.05 mg/kg. The acute oral LD50 ranges from 0.3 to 2.3 mg/kg in the rat and from 3.0 to 7.5 mg/kg in the dog. It was found to be 14.7 mg/kg for cats and 150 mg/kg for pigs. In mice and rabbits, the oral LD50 is 340 and 35 mg/kg, respectively. For the mallard duck, it is 3158 mg/kg. The LD50 associated with 14 daily oral doses in rats was 0.1 mg/kg (Correll et al., 1952). Acute percutaneous LD50 for rats is less than 200 mg/kg. In a 21-day subchronic percutaneous study in rabbits, the no-effect level was 0.1 mg/kg daily. Diphacinone is neither a skin and eye irritant nor a skin sensitizer. An acute inhalation of diphacinone dust in the rat has shown an LC50 of less than 2000 mg/m3 of air. Diphacinone is not mutagenic in the Ames test. Sprague-Dawley rats were fed for 21 days on a diet containing 1, 2, or 4 ppm diphacinone. All animals in the 2- and 4-ppm groups died and postmortem examination revealed massive internal hemorrhages. On day 21, the prothrombin clotting time of the animals in the 1-ppm group was not affected. A second study was conducted in which the rats were fed for 90 days on a diet containing 0.0313, 0.625, 0.125, 0.25, or 0.5 ppm of diphacinone, or approximately 0.002, 0.003, 0.006, 0.013, and 0.025 mg/kg/day. One male in the 0.25-ppm group died on day 17 of treatment and another male in the 0.0625-ppm group on day 20 from a subdural hemorrhage. However, the mean prothrombin

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clotting times of the animals surviving the treatment period were not affected. The only parameter that showed some variation was the fibrinogen level, which was lower in the 0.5-ppm group (Elias and Johns, 1981). The most interesting aspect of the toxicity of diphacinone involves species difference. The oral LD50 of the compound for vampire bats (Desmodus rotundus) was 0.91 mg/kg, whereas a dosage of 5 mg/kg produced no sign of illness in cattle (Elias et al., 1978). The blood of beef cattle given a single intraruminal injection of the compound at a rate of 1 mg/kg became toxic to these bats and remained toxic for 3 days without harming the cattle. As indicated by examination before and 2 weeks after treatment, cattle dosed in this way on three ranches in Mexico experienced a 93% reduction in vampire bat bites. Bioassays of milk and liver indicated that there was no residue problem (Thompson et al., 1972). Residue studies indicated that people may safely eat meat, including liver and kidney, from treated cattle (Bullard et al., 1976). In 1996, the California Environmental Protection Agency (CalEPA) updated its toxicological database on diphacinone and has published summaries of some useful studies (California Environmental Protection Agency, 1996). In a teratogenicity study, diphacinone technical (purity 98%) administered by gavage to pregnant rats from day 6 through day 15 of gestation at nominal concentrations in corn oil of 0.10, 0.025, or 0.075 mg/kg bw/day did not induce any malformation including at the high dose. Some equivocal embryotoxic effects (early resorption and postimplantation losses) were noted at this dose level. The actual maternal NOAEL was 0.0195 mg/kg bw/day (nominal: 0.025 mg/kg bw/day). A 14-day oral toxicity study is also briefly summarized in the CalEPA document. Sprague-Dawley rats of both sexes received 14 consecutive daily doses of technical diphacinone at concentrations of 0.025, 0.040, 0.085, or 0.175 mg/kg bw/day in corn oil. Unfortunately, no histopathology was performed. Deaths occurred at 0.175 mg/kg bw/day and the NOEL was determined to be 0.04 mg/kg bw/day for coagulopathy and at 0.085 mg/kg bw/day for general toxicity and lethality. The same document briefly summarized three genotoxicity studies: gene mutation in the Salmonella typhimurium test, mammalian cell forward mutation assay (CHO/ HGPRT assay), and DNA damage in the mouse bone arrow micronucleus test. All three tests gave negative results. (a) Absorption, Distribution, Metabolism, and Excretion When 14C-labeled diphacinone was administered orally to mice, radioactivity reached its highest levels in the liver and lungs. The concentration in the liver reached its maximum in 7.5 and 3.0 h in males and females, respectively (Cahill and Crowder, 1979). In another study, rats and mice were orally administered 14C-labeled diphacinone at dosages of 0.2 and

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1.5 mg/kg. In rats, about 70% of the dose was excreted in the feces and 10% in urine in 8 days. The same elimination pattern was observed in mice. Eight days after the administration of the compound in rats and 4 days in mice, the liver had the highest level of residues, but kidneys and lung also contained significant levels of residues; brain, fat, and muscles had the lowest levels. Diphacinone is not extensively metabolized in rats, less than 1% of the dose being expired as CO2. The metabolism pattern in rats involved mainly hydroxylation and conjugation reactions (Yu et al., 1982). An analytical method has been developed to determine and quantify residues of diphacinone in the tissues of invertebrates. It is a reversed-phase ion-pair liquid chromatography with a solid-phase extraction clean-up and UV detection (Primus et al., 2006).

with jaundice, nephropathy with acute renal tubular necrosis, severe exfoliative dermatitis, and massive generalized edema (American Medical Association, 1977). The drug ceased to be listed in the AMA Drug Evaluations of 1980. Diphacinone, like other hydroxycoumarin rodenticides, probably crosses the skin barrier. This is illustrated by the case reported by Spiller et al. (2003) of a professional exposure of an 18-year-old male pest control employee who spilled a concentrated liquid preparation of 0.106% chlorophacinone in his boot. The man did not remove his boot for 6 or 8 h. Seven days later, he was hospitalized with flank pain, hematuria, and epistaxis. Vitamin K treatment was undertaken (without associated blood products because of religious beliefs). Diphacinone was detected in the serum drawn 60 days postexposure at which time his prothrombin time was back to normal.

(b) Mode of Action Diphacinone inhibits the K-enzyme complex (liver-synthesized coagulation proteins: factors II, VII, and X), and this inhibition phase lasts approximately 30 days in dogs, as opposed to the relatively short effect of warfarin (Mount and Feldman, 1983). The prolonged action of diphacinone may be due to protein binding in the liver or low excretion rate or a combination of both factors.

(b) Treatment of Poisoning Treatment is the same as for warfarin, but based on experimental data from animals it would seem advisable to increase the dose of vitamin K as well as the duration of the corrective treatment.

(c) Treatment of Poisoning in Animals In dogs, the results of the usual vitamin K oral therapy after diphacinone poisoning were poor. Dogs treated with a sufficient quantity of vitamin K relapsed fatally several days after the corrective treatment stopped. The response to treatment seems to vary according to the amount of exposure to the rodenticide. The recommended therapeutic dose of vitamin K is 5 mg/kg of body weight in subcutaneous injections for at least 5 consecutive days (Mount and Feldman, 1983).

100.5.5.3  Toxicity to Humans (a) Therapeutic Use Diphacinone has been used as a therapeutic agent because it has a relatively long duration of action. Its half-life in humans is 15–20 days. A single oral dose of diphacinone of 4 mg/person produced a clearly detectable reduction of prothrombin about 14 h after ingestion and slightly more reduction a day later, with recovery to normal by the 3rd day. A smaller, uncertain reduction was produced by 2 mg/person. A single 20-mg dose caused hypoprothrombinemia that was definite in 14 h, marked in 48 h, and persisted from 6 to 10 days. The recommended initial dose for therapy was 20 mg followed by daily doses of 2–4 mg (Field et al., 1952). The drug was in use until a few years ago. Although there were no adverse effects except occasional nausea and not unexpected hemorrhagic complications at high dosage levels, caution was advised because of its close relation to phenidione, which has caused agranulocytosis, hepatitis

100.5.6  Brodifacoum 100.5.6.1  Identity, Properties, and Uses (a) Chemical name 3-[3-(4’-Bromo-[1,1’-biphenyl]-4-yl)-1,2,3,4-tetrahydro-1naphthylenyl]-4-hydroxy-2H-1-benz-o-pyran-2-one is the chemical name. (b) Structure See Figure 100.4. (c) Synonyms Brodifacoum is the approved common name (ISO-BIS). Trade names for the formulated material include Ratak, Volak, and Talon. The code names are WBA 8119 and PP 581. The CAS registry number is 56073-10-0. (d) Physical and Chemical Properties The empirical formula for brodifacoum is C31H23BrO3 and the molecular weight is 523.4. It is an off-white to fawn-colored odorless powder with a melting point of 228–232°C. It is of very low solubility in water (less than 10 mg/l at 20°C and pH 7). Brodifacoum is slightly soluble in alcohols and benzene and soluble in acetone. It is stable at room temperature. It has a very low vapor pressure of less than 1.33  107 kPa (1  106 mmHg) at 25°C. (e) History, Formulation, and Uses The rodenticidal properties of brodifacoum were described in 1976. It is an indirect anticoagulant active against rats and mice including strains resistant to warfarin and other anticoagulants (Rennison and Hadler, 1975). It is also used

Chapter | 100  Rodenticides

O

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O

OH

Difenacoum Br O

O

OH

Brodifacoum Figure 100.4   Chemical structure of two synthetic anticoagulant compounds: difenacoum and brodifacoum. 

Table 100.5  Single-Dose LD50 for Brodifacoum Species

Route

LD50 (mg/kg)

Reference

Rat, M

Oral

0.27

Tomlin (1994)

Rat

Dermal

50.00

Tomlin (1994)

Mouse, M

Oral

0.40

Tomlin (1994)

Guinea pig, F

Oral

0.28

Tomlin (1994)

Rabbit, M

Oral

0.30

Tomlin (1994)

Dog

Oral

0.25–1.0

Tomlin (1994)

Cat

Oral

0.25

Tomlin (1994)

to control other wild rodents. Brodifacoum is formulated as ready-to-use baits of low concentration (20 and 50 mg/kg of bait). A single ingestion is usually sufficient to kill.

100.5.6.2  Toxicity to Laboratory Animals Brodifacoum is extremely toxic to a number of mammalian species. The oral and dermal LD50 values of the technical material are given in Table 100.5. In chicken, the oral LD50 is reported to be 4.5 mg/kg and in the mallard duck it is 2.0 mg/kg. In a 42-day feeding study in rats, a concentration of 0.1 ppm did not induce any adverse effect (Worthing and Walker, 1983).

The prolonged effects of brodifacoum after a single oral administration (brodifacoum: 0.2 mg/kg bw) to male Wistar rats have been investigated (Mosterd and Thijssen, 1991). Various parameters related to blood clotting activity and liver parameters of the vitamin K cycle were regularly investigated over a period of 30 days after dosing. The main observations made were the maximum effect on blood clotting activity was seen on day 1, which returned to normal on day 7 posttreatment; liver microsomal vitamin K epoxide reductase activity was maximally reduced (down to 10% of control) also on day 1 and recovered to approximately 40% on day 15 and then remained at this level; the persistent inhibition of the vitamin K cycle was also verified in vivo; following intravenous vitamin K administration of 10 mg/kg bw on day 30, the brodifacoum-treated rats accumulated vitamin K epoxide in the liver; brodifacoum rapidly accumulated in the liver until saturation of the microsomal binding site, preventing its elimination from the liver; and the liver content on day 30 was not different from day 7. In 2005, the California Environmental Protection Agency updated its toxicological database on brodifacoum and published summaries of a number of useful studies (California Environmental Protection Agency, 2005). No teratogenic effects were detected in pregnant Wistar rats orally treated with brodifacoum technical (92.5%) from day 6 through day 15 inclusively at doses of 0.001, 0.01, or 0.02 mg/kg bw/day. The maternal NOEL was 0.001 mg/kg bw/day based on uterine hemorrhages at higher dosages. The developmental NOEL is  0.02 mg/kg bw/day. In pregnant rabbits orally exposed to brodifacoum technical (92.5%) from day 6 through day 18 of gestation at doses of 0.001, 0.002, or 0.005 mg/kg bw/day, no malformations were observed but a high incidence of maternal mortality due to hemorrhages was seen in the high-dose group. The maternal NOEL was determined to be 0.002 mg/kg bw/day and the developmental NOEL 0.005 mg/kg bw/day. An Ames test and a mouse bone marrow micronucleus test are also summarized in the CalEPA document, both with negative results (California Environmental Protection Agency, 2005).

100.5.6.3  Toxicity to Nontarget Animals Several cases of poisoning in domestic animals have been reported. One day after ingestion of brodifacoum containing bait was observed, a 17-kg cocker spaniel developed depression and icterus accompanied by accelerated pulse and rapid and labored respiration. Despite supportive therapy, the dog died the same day. The autopsy confirmed the icterus and showed approximately 1 l of unclotted blood in the thoracic cavity and 100 ml in the pericardiac sac. Numerous hemorrhagic areas were seen in the serous membranes (Stowe et al., 1983). A 4-year-old cross-bred bitch was noticed to be depressed and weak the day following

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the laying of Talon bait (0.005% brodifacoum); she was found dead in her kennel the next morning. The actual dose of brodifacoum ingested was unknown but it was estimated that the maximum quantity of bait eaten could have been 900 g, resulting in an intake of 45 mg of active ingredient. At autopsy, the thoracic cavity contained approximately 1.8 l of unclotted blood and a single clot was adherent to the base of the heart and to the aorta. Subcutaneous bruising was present on the rib cage. Brodifacoum was found in the liver at a concentration of 0.8 mg/kg (McSporran and Phillips, 1983). Eason and Spurr (1995a) reviewed the toxicity and impact of brodifacoum on nontarget wildlife in New Zealand. They conclude that brodifacoum has the potential to cause primary and secondary poisonings of nontarget species but that the adverse effects depend more on how and where baits are used and on the feeding behavior of the nontarget species than on the susceptibility of individual species. However, the risks of brodifacoum to nontarget birds and mammals in New Zealand were reviewed by the same authors and they concluded that the risks to nontarget birds and other wildlife are determined to a significant extent by species’ intrinsic susceptibility and the toxicokinetics of the compounds used (Eason et al., 2001). Robben et al. (1998) report cases of poisoning in 21 dogs in which brodifacoum, difethialone, and difenacoum were detected by HPLC in the plasma of 13, three, and two dogs, respectively. The initial plasma concentrations ranged from below the limit of detection (10 ng/l) to 851 ng/l. No anticoagulant substance could be detected in three dogs even though these animals were exhibiting characteristic signs of coagulation disturbances and positive response to vitamin K administration. In seven dogs, the half-life of brodifacoum ranged from 0.9 to 4.7 days (median, 2.4) with a mean residence time (MRT) of 1.9 to 3.7 days (median 2.8). In two dogs, the individual half-life of difethialone was 2.2 and 3.2 days and the MRT was 2.3 and 2.8 days. Borst and Counotte (2002) report an unusual case of poisoning occurring in a nontarget species. The episode took place in a zoo where the first two offspring of a pair of turkey vultures (Cathartes aura) died of brodifacoum poisoning. The adult birds fed rodenticide-killed mice to their offspring. The authors make reference to previous reports of cases of small carnivorous birds found dead after eating poisoned mice. The acute oral toxicity of brodifacoum was examined in sheep. An LD50 of 11 mg/kg was considered a good estimate (Godfrey et al., 1985). A white-winged duck (Cairina sculata) in a zoological setting was found with bilateral epistaxis and anemia. Brodifacoum was detected in the blood at a level of 0.002 ppm. The bird was treated with injectable and oral vitamin K and transfused with 40 ml of whole blood and fully recovered (James et al., 1995).

Hayes’ Handbook of Pesticide Toxicology

Cases of poisoning have been described in other species (Ayala et al., 2007; Eason and Spurr, 1995a,b). (a) Absorption, Distribution, Metabolism, and Excretion Brodifacoum is absorbed through the gastrointestinal tract. When orally administered to male Sprague-Dawley rats at doses ranging from 0.1 to 0.33 mg/kg, brodifacoum exhibited a remarkably steep dose–response curve; 0.1 mg/kg failed to show an effect on the plasma prothrombin level within 24 h, whereas 0.2 mg/kg reduced the prothrombin complex activity to 7% of normal values and 0.33 mg/kg reduced it to 4% of normal. Concentrations in the liver were rapidly established and remained relatively constant for at least 96 h. The mean liver/serum concentration ratio is approximately 20. Disappearance from serum is slow, with a half-life of 156 h or even more. The slow disappearance from the plasma and liver and the large liver/serum ratio probably contribute to the higher toxicity of brodifacoum compared to warfarin. These particular features may also explain the efficacy of brodifacoum against warfarin-resistant rats (Bachmann and Sullivan, 1983). Six weeks after intravenous administration of a single 1-mg/kg dose of brodifacoum to male New Zealand White rabbits, the prothrombin complex activity was still lower than 30% of normal (in the early part of the study, subcutaneous injections of vitamin K were given to prevent lethal hemorrhage). In the same study (Park and Leck, 1982), it was shown that in the rabbit, the maximal antagonism of vitamin K by warfarin was produced by a dose of 63 mg/kg, whereas a similar result was obtained with only 1 mg/kg brodifacoum. It was shown that, in warfarin-resistant and warfarin-sensitive rats, brodifacoum produced the same rate of degradation of prothrombin complex activity as warfarin and significantly reduced the activity of clotting factors II, VII, and X without affecting factor V. It was also demonstrated that brodifacoum has the same mechanism of action as warfarin: reduction of vitamin K-dependent clotting factor synthesis by interruption of the vitamin K epoxide cycle (Leck and Park, 1981). In mongrel dogs, the elimination of brodifacoum follows a classical experimental decay with a distributive half-life of 1.4 days and an elimination half-life phase of 8.7 days (Murphy et al., 1985). Brodifacoum was fed to four dogs for 3 consecutive days, producing a cumulative dose of 1.1 mg/kg body weight. Serum brodifacoum concentrations were monitored. Inappetence and hemorrhagic tendencies were exhibited by day 5 after rodenticide exposure. One-stage prothrombin time, APTT, and ACT were 25% greater than time zero values at 24, 24, and 72 h after dosing, respectively. All laboratory parameters returned to normal within 48 h of initiating vitamin K therapy (0.83 mg/kg orally for 5 days). Serum brodifacoum concentrations were highest (1065–1215 ng/ml) during the 3 days after dosing and were

Chapter | 100  Rodenticides

detectable (3.0–7.5 ng/ml) until day 24 after brodifacoum exposure. A mean brodifacoum elimination half-life of 6  4 days was observed (Woody et al., 1992). The case of brodifacoum toxicosis in two Great Pyrenean puppies reported by Munday and Thompson (2003) clearly indicates that brodifacoum readily crosses the placenta. Eight out of a litter of 13 puppies were either born dead or died within 48 h of birth. Three puppies that died shortly after whelping were necropsied. Two of them had hemorrhage in the thoracic and peritoneal cavities, intestinal serosa, and meninges. Brodifacoum was detected in the livers of these two puppies. The dam did not have clinical signs of coagulopathy before or after whelping and was not exposed to brodifacoum for at least 4 weeks before whelping. In a survey of residues of brodifacoum and other anticoagulant rodenticides in target and nontarget species, Spurr et al. (2005) detected residues of brodifacoum in the livers of 234 mammals from eight species and in two birds from two species captured alive. The highest brodifacoum concentrations in mammalian livers were recorded during the period brodifacoum baits were used but residues were also present in the livers of some individuals at least 24 months after brodifacoum use in the area stopped. According to the authors, these residues may have persisted in animals surviving brodifacoum use in the area and/or been transported into the area by animals moving to and from adjacent villages and farms where brodifacoum use continued. (b) Factors Influencing Toxicity Pretreatment of rats daily by intraperitoneal injection of phenobarbital at 80 mg/kg for 2 consecutive days followed by a single administration of brodifacoum at 0.2 mg/kg by stomach tube reduced the anticoagulant effects, although the reduction was less marked than in the case of warfarin (Bachmann and Sullivan, 1983). It is also known that a very large number of drugs of different chemical structures can interact with coumarin anticoagulant therapy in humans (Koch-Weser and Sellers, 1971). (c) Pathology Necropsies of poisoned dogs have shown that, in addition to large collections of unclotted blood, a number of lesions were also present, including bile stasis with large amounts of brown pigment accumulated in the portal triads and in the macrophages of the periportal regions and congestion of the spleen with accumulation of golden brown pigment, believed to be hemosiderin, in the red pulp (Stowe et al., 1983). (d) Treatment of Poisoning in Animals Treatment should be as for other anticoagulant rodenticides: vitamin K at dosage of 2.5–5.0 mg/kg and transfusion of fresh whole blood. Because of the long-lasting effect of brodifacoum, vitamin K therapy must be continued for at least 2–3 weeks.

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100.5.6.4  Toxicity to Humans Brodifacoum has not been used therapeutically in humans. Cases of attempted suicide have been reported. A 31-yearold mentally disturbed woman ingested over a 2-day period approximately thirty 50-g packages of Talon (approximately 75 mg of brodifacoum). Two days later she was brought to the hospital’s psychiatric unit, without any physical signs or symptoms. The routine laboratory tests showed a prothrombin time of 72 s (control, 12 s) and an activated partial thrombo­plastin time greater than 100 s (normal, 25–35 s). In spite of prolonged administration of large amounts of vitamin K and repeated infusion of fresh frozen plasma, the depression of the prothrombin complex activity persisted for more than 45 days after the ingestion (Lipton and Klass, 1984). Jones et al. (1984) reported a similar case in a 17year-old boy who attempted suicide by ingesting approximately 7.5 mg (0.12 mg/kg) brodifacoum. He was first seen for a gross hematuria, rapidly followed by epistaxis and gum bleeding. The prothrombin time and the activated partial thromboplastin time were considerably prolonged. The levels of plasma clotting factors II, VII, IX, and X were decreased. Factor V was normal. Vitamin K and plasma therapy were instituted and had to be continued for 55 days until the patient’s coagulation remained normal and stable. The average fatal dose for an adult 60-kg man was recently estimated to be 15 mg without treatment (Anonymous, 2007b). (a) Dosage Response In the two cases of poisoning reported above, the total doses ingested were in one case 75 mg and in the second one 7.5 mg and the effects on the clotting factors were maximum in both cases, the clinical signs being almost absent in the woman who absorbed 75 mg of brodifacoum (although it was ingested over a 48-h period). Therefore, it would seem that, above a certain threshold, the response is maximum. This was also shown in rabbits given 1 and 10 mg/kg brodifacoum (Park and Leck, 1982). Murphy et al. (1985) showed that, in dogs, serum concentrations below 12 mg/ml caused no measurable coagulopathic effects after cessation of vitamin K therapy. Plasma brodifacoum concentration was compared to prothrombin levels over time in a case of brodifacoum poisoning. Brodifacoum was eliminated according to a twocompartment model, with an initial half-life of 0.75 days and a terminal half-life of 24.2 days. On admission, the brodifacoum level was 731 g/l and the patient suffered severe urinary tract hemorrhage, requiring transfusion of blood products. Persistently increased prothrombin times necessitated treatment with phytonadione up to 80 mg/day for 4 months, until the brodifacoum level reached 10 g/l (Hollinger and Pastoor, 1993). The plasma concentration, plasma half-life, and mean retention time of brodifacoum (among other anticoagulant

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rodenticides) were determined in dogs in which preliminary diagnosis of anticoagulant poisoning had been made. Analysis was performed with HPLC on the plasma. In seven dogs, the estimated half-time of brodifacoum ranged from 0.9 to 4.7 (median, 2.4) days with a mean retention time of 1.9 to 3.7 (median, 2.8 days) (Robben et al., 1998). The case of a voluntary ingestion of brodifacoum by a 39-year-old man was reported by Sheen et al. (1994). His prothrombin and partial thromboplastin times were, respectively, 150 and 113s. Treatment with a daily dose of 200 mg of phytonadione for 5 months corrected the coagulopathy with no side effects occurring. Prolonged follow-up and vitamin K treatment were necessary in a child with spontaneous hemorrhage from her nose, mouth, and urinary tract following accidental ingestion of brodifacoum (Travis et al., 1993). The case of a woman who survived brodifacoum poisoning is interesting as it provides information regarding the characteristic long half-life of the compound and the necessary long period of treatment. The patient was hospitalized with severe coagulopathy. Serum brodifacoum levels were determined by HPLC over 7 months. Five days after admission, the level was 1302 ng/ml, slowly decreasing over a period of 209 days when it became undetectable. This follow-up demonstrated first-order kinetics and a half-life of 56 days for brodifacoum (Olmos and Lopez, 2007). Analytical methods are available to detect and to quantify brodifacoum and other coumarin derivative rodenticides in various substrates such as foodstuffs and stomach content of animals suspected of poisoning (Adamowicz and Kala, 2005) as well as in human serum (Kuijpers et al., 1995). Multi-compound methods are available for the simultaneous determination and quantification of indandiones and hydroxycoumarins in blood, serum and liver (Chalermchaitkit et al., 1993; Vandenbroucke et al., 2008).

100.5.7  Chlorophacinone 100.5.7.1  Identity, Properties, and Uses (a) Chemical Name 2-[4-(Chlorophenyl)phenylacetyl]-1H-indene-1,3(2H)dione is the chemical name. (b) Structure See Figure 100.3. (c) Synonyms Chlorophacinone is the approved common name (BSIISO). Trade names for the formulated products include Caid, Liphadione, Raviac, Drat, Quick, Lepit, Rozol, and Saviac. The CAS registry number is 3691-35-8.

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(d) Physical and Chemical Properties Chlorophacinone has the empirical formula C23H15ClO3 and the molecular weight of 364.8. It forms a yellow crystalline solid with a melting point of 140°C. It is slightly soluble in water (100 mg/l at 20°C) and soluble in acetone, ethanol, and methanol. It is stable under normal storage conditions and noncorrosive. (e) History, Formulation, and Uses Chlorophacinone is an anticoagulant rodenticide used to control rats, mice, voles, and other wild rodents. It is formulated as ready-to-use baits based on whole, cracked, or milled grain at concentrations of active ingredient ranging from 0.005 to 0.25%. It can also be used as a tracking powder. An oil concentrate is also available.

100.5.7.2  Toxicity to Laboratory Animals The acute oral LD50 is reported to be 2 mg/kg in the rat, 1 mg/kg in the mouse, and 50 mg/kg in the rabbit. In the duck, the oral LD50 is 100 mg/kg. Chlorophacinone is of low acute toxicity to wild birds (LD50 of 430 mg/kg). The acute dermal LD50 in the rabbit is 200 mg/kg (Sax, 1984). It is absorbed through the skin of the rabbit; a solution of 5 mg in 2 ml of liquid paraffin applied to 100 cm2 of shaved skin of a rabbit caused a slight reduction of prothrombin (Worthing and Walker, 1983). Administration of 15 daily doses of 2.25 mg to gray partridges produced no detectable ill effects. Chlorophacinone is not an eye or skin irritant. No published subchronic toxicity study seems to be available except one briefly summarized by the California Environmental Protection Agency (2002) but found to be unacceptable because of poor reporting quality. This study conducted on Sprague-Dawley rats orally exposed for 3 months at daily doses between 0.5 (terminated at 77 days) and 160 g/kg bw/day provided some information of the toxic effects of chlorophacinone. Animals in dose groups 40 g/kg bw/day and above had excessive and rapid mortality and signs of characteristic anticoagulant intoxication. Hemorrhages in various organs were found in almost all animals at 20 g/kg bw/day. The lowest observed effective level (LOEL) was 10 g/kg bw/day. The same CalEPA document briefly reports teratogenicity studies in rats and rabbits. In pregnant CD7 SpragueDawley rats receiving pure chlorophacinone by gavage at dose levels of 12.5, 25, 50, or 100 g/kg bw/day from gestation day 6 through 15 inclusively, mortality was 72% in dams of the high-dose group (100 g/kg bw/day). The identified cause of death resulted from anticoagulation disturbances. The maternal NOEL was determined to be 50 g/kg bw/day. In the fetuses, the majority of visceral malformations were unilateral or bilateral hydroureter and the upward trend was dose-related. However, since hydroureter is a relatively common finding in this strain of SpragueDawley rats and is usually reversible in the offspring and the litter incidence was not significant by pair-wise

Chapter | 100  Rodenticides

comparison for total ureter findings, it is difficult to decide whether or not this observation is related to an effect of chlorophacinone itself. The developmental NOEL was considered to be 12.5 g/kg bw/day. In NZW pregnant rabbits, chlorophacinone (purity of 101%) was administered by gavage (in corn oil) at doses of 5, 10, 25, or 75 g/kg/day on gestational days 7 through 19. Mortality was 100% and 81% for dose groups 75 and 25 g/kg/day, respectively. Clinical observations and necropsy findings indicated anticoagulant activity (external and internal hemorrhaging). Maternal NOEL was 10 g/kg/day (mortality and clinical signs) and the developmental NOEL was 10 g/kg/day (at 25 g/kg/day, there were two surviving litters and, although there were no effects in these litters, the number was too low for a definitive evaluation). There were no treatment-related developmental findings and the NOEL for teratogenicity was determined to be the highest administered dose (75 g/kg bw/day). Three valid genotoxicity tests are reported in the CalEPA summary of toxicological data document: an Ames test, an HGPRT test in CHO cells, and a human lymphoma cell test. All three gave negative results.

100.5.7.3  Toxicity to Laboratory Animals A series of poisonings occurring in lambs is reported by Del Piero and Poppenga (2006). Eleven lambs, approximately 1–2 months old, suddenly developed epistaxis, respiratory distress, and facial and cervical swellings. The affected animals died within 2 h of onset of clinical signs. Two lambs were subjected to complete postmortem examination. At necropsy, gross lesions were characteristic of profuse bleeding. Histologically, hepatocellular centrolobular necrosis was observed. Chlorophacinone was detected in the livers at 0.58 and 0.50 ppm. The source of exposure was old baits placed between the walls of the building housing the ewes and lambs. The British Veterinary Laboratories Agency conduct regular surveys of animal toxicosis and publishes the results (Sharpe and Livesey, 2005). Such surveys are also performed and published by the Scottish Department of Environment and Rural Affairs (Hunter et al., 2005). (a) Absorption, Distribution, Metabolism, and Excretion Chlorophacinone is absorbed through the gastrointestinal tract. After oral administration, 90% is eliminated in the feces within 48 h in the form of metabolites (Hartley and Kidd, 1983). (b) Mode of Action Chlorophacinone is an anticoagulant agent depressing hepatic synthesis of prothrombin and clotting factors VII, IX, and X. Direct damage to capillary permeability occurs concurrently. The ultimate effect of these actions is to

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induce widespread internal hemorrhage. In addition, chloro­ phacinone is an uncoupler of oxidative phosphorylation. Unlike the coumarin derivatives, chlorophacinone may cause symptoms and signs of neurologic and cardiopulmonary injury in laboratory rats, which often lead to death before hemorrhage occurs. Chlorophacinone is characterized by its long-lasting depressive action on coagulation.

100.5.7.4  Toxicity to Humans Chlorophacinone has not been used therapeutically. Several reports of poisoning cases, either accidental or deliberate, have been published. A 37-year-old woman ingested about 250 ml of a 0.25% concentrate formulation (about 625 mg of chlorophacinone). Despite intensive therapy with vitamin K (phytomenadione, a natural form of vitamin K), the anticoagulant effect of chlorophacinone persisted for at least 45 days. Interestingly, this episode showed that the synthetic analogue of vitamin K was ineffective (Murdoch, 1983). Another report concerns a 28-year-old man who ingested an unknown amount of chlorophacinone-based rodenticide. Again, the most striking feature in this case was the unusually prolonged and severe anticoagulant effect, even under adequate therapy; it required 4 weeks for prothrombin level to return to normal (Dusein et al., 1984). Vogel et al. (1998) reported the case of an 18-yearold woman who was admitted to the hospital 3 days after ingesting approximately 100 mg of chlorophacinone. Under high-dose vitamin K therapy, her prothrombin time (PT) was normalized but it increased again when the treatment was withdrawn. Prolonged administration of vitamin K for 7 weeks was necessary before PT became normal. (a) Treatment of Poisoning Intoxication by chlorophacinone is treated by massive and prolonged administration of natural vitamin K. However, the main feature of chlorophacinone poisoning is its long-lasting effect on the coagulation parameters (Burucoa et al., 1989). Some reports show that monitoring the chlorophacinone plasma concentrations in intoxicated patients may provide useful information on toxicological management. A 33-year-old man who voluntary ingested 750 ml of a liquid formulation corresponding to 1875 mg of chlorophacinone had a normal prothrombin index (PI) when he was admitted to the hospital 8 h after ingestion. The maximum plasma level was then 27.6 mg/ml. To search for an optimum therapeutic scheme based on phytomenadione (vitamin K) associated with phenobarbital as a cytochrome P450 inducer, PI, and chlorophacinone plasma levels were monitored for 17 days. Under phenobarbital 200 mg/day, chlorophacinone exhibited an elimination half-life of 3.27 days, shorter than previously reported. On the basis of their experience with this case, the authors note that monitoring the chlorophacinone concentrations may provide a

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better estimation of the necessary duration of hospitalization and treatment (Lagrange et al., 1999). This approach is also recommended by Bruno et al. (2000) based on their own experience with one of their patients who voluntarily ingested a large quantity of brodifacoum. This question has been recently reviewed (Watt et al., 2005).

100.5.8  Difenacoum 100.5.8.1  Identity, Properties, and Uses (a) Chemical Name 3-[3-(1,1’-Biphenyl)4-yl-1,2,3,4-tetrahydro-1-naphthalenyl]4-hydroxy-2H-1-benzopyran-2-one is the chemical name. (b) Structure See Figure 100.4. (c) Synonyms Difenacoum is the approved common name. Trade names include Neoxorexa and Ratak. The CAS registry number is 56073-07-5. (d) Physical and Chemical Properties Difenacoum has the empirical formula C31H24O3 and a molecular weight of 444.5. It is an off-white powder with a melting point of 215–219°C. It is slightly soluble in water (less than 10 mg/l at pH 7) and soluble in organic solvents (50 g/l in acetone and chloroform and 600 mg/l in benzene). (e) History, Formulations, and Uses The rodenticidal properties of difenacoum were first described in 1975. It is formulated as a 1-g/kg concentrate and as a ready-to-use bait containing 50 mg of active ingredient per kilogram of bait. It is an indirect anticoagulant, more potent than the early compounds. It is used to control rats and mice resistant to other anticoagulants with varying degrees of activity.

100.5.8.2  Toxicity to Laboratory Animals The oral LD50 is 1.8 mg/kg in male rats, 0.8 mg/kg in male mice, and 50 mg/kg in female guinea pigs. The LD50 value for oral administration in pigs is reported to be above 80 mg/kg and it is 100 mg/kg in cats. The acute dermal LD50 is 50 mg/kg in rats and 1000 mg/kg in rabbits. The cumulative oral LD50 in male rats over a 5-day period is 0.16 mg/kg/day. Winn et al. (1987) investigated the actions of difena­ coum in male and female rats and mice. The animals were administered a single oral dose of 0.5 mg/kg bw difenacoum. This dose killed 50% of male mice with 9 days of administration, whereas no female mice died during the

Hayes’ Handbook of Pesticide Toxicology

same time. Baseline values of the prothrombin complex activities (PCA) of male and female rats were not significantly different (0.4 mg/kg bw intraperitoneally). Over the first 24 h following difenacoum administration, a monoexponential fall in the PCA was seen in both sexes. However, 6, 12, and 24 h after treatment, the PCA in male rats was significantly lower than in female rats. The PCA started to recover over the subsequent 48 h in both sexes. The difference between the onset of action of difenacoum in males and females did not appear to be due to a greater rate of elimination of the compound in females because the plasma concentrations 24 h after administration were the same in both sexes. The initial predosing concentrations of vitamin K in the liver of male and female rats were 35.1  18.6 and 29.4  5.4 ng/g, respectively liver, but 24 h after difenacoum administration, the vitamin K levels were either very low or undetectable in all rats. A series of toxicological studies have been recently evaluated and summarized by the California Environmental Protection Agency (CalEPA, 2006). In a teratology study, Wistar rats received single oral daily doses of 0.01, 0.03, or 0.09 mg/kg bw of technical difenacoum. Hemorrhages were found in dams in the highdose group. No malformations or embryotoxicity were recorded at any dose levels. The developmental NOEL was 0.09 mg/kg bw/day and the maternal NOEL was 0.03 mg/kg bw/day. In NZW rabbits treated with single oral doses of 0.001, 0.005, or 0.015 mg/kg bw/day, technical difenacoum induced severe disturbances in the coagulation parameters in the high-dose animals without clinical signs (but the animals were killed for humane reasons before bleeding occurred). The maternal NOEL was 0.005 mg/kg bw/day and the developmental NOEL was 0.015 mg/kg bw/day as no developmental effects were noted at the high-dose level. A series of nine genotoxicity studies were all negative.

100.5.8.3  Toxicity to Nontarget Animals Difenacoum and brodifacoum have been suspected of being responsible for secondary toxicity in barn owls feeding on rodents poisoned by these second-generation anticoagulants, a phenomenon that was not seen with warfarin baits (Wenz, 1984). This aspect of the particular toxicity of second-generation anticoagulant rodenticides has been investigated on three compounds including difenacoum (in addition to brodifacoum and flocoumafen). Those compounds were separately fed to barn owls (Tyto alba, Scop.) for 15 days via rodenticide-fed mice to simulate the actual route of exposure. The birds survived a cumulative dose of each rodenticide of at least 1.9 mg/kg bw each day for 15 days. Residues were analyzed in the owls and the results confirmed that the liver was the organ retaining the largest quantity of ingested rodenticide (Gray et al., 1999).

Chapter | 100  Rodenticides

Similarly, in a survey done in the United Kingdom by Shore et al. (1999), at least 25–35% of individuals in populations of small mammal predators are secondarily exposed to second-generation anticoagulants in Britain. However, the authors indicate that there is little understanding of the frequency with which this occurs in mammals or its importance compared to other causes of mortality and that further research is needed to assess the magnitude and frequency of rodenticide intake by small mammal predators and to determine the toxic effects of this exposure. (a) Biochemical Effects Like brodifacoum and warfarin, difenacoum was shown to inhibit K-dependent steps in the synthesis of clotting factors II, VII, IX, and X, and it is suspected that coumarin anticoagulants block the vitamin K epoxide cycle by inhibiting the vitamin K epoxide reductase. The latter is confirmed by the observation that difenacoum and brodifacoum produce an accumulation of tritiated vitamin K epoxide in rats and rabbits administered tritiated vitamin K (Park and Leck, 1982). Like brodifacoum, difenacoum has a much longer duration of action than warfarin. Parmar et al. (1987), cited in WHO Environmental Health Criteria 175, page 51 (1995), have reported a rapid elimination phase from the liver lasting from day 2 to day 8 after dosing and a slower terminal phase with elimination half-lives of 130 days for brodifacoum and of 120 days for difenacoum (compared to 170 days for bromadiolone). However, the determination of the half-life values has been questioned on the basis of methodology (for example, the number of experimental points); therefore, the study duration may also influence the halflife value (Lechevin and Vigie, 1992). Parmar et al. (1987) has shown that difenacoum is largely metabolized in rats and mostly eliminated as metabolites, whereas brodifacoum and coumatetralyl are eliminated as unchanged parent compounds. (b) Treatment of Poisoning in Animals Prolonged administration of vitamin K over several weeks is the treatment of choice. Since the effect of vitamin K is usually delayed, because it only permits the formation of new prothrombin, initial treatment with transfusion of fresh-frozen plasma or a small quantity of matched fresh blood is recommended in order to provide enough prothrombin to prevent further hemorrhage (Barlow et al., 1982; Park and Leck, 1982).

100.5.8.4  Toxicity to Humans A case of an attempted suicide in a 17-year-old girl is reported in the literature. She was admitted to hospital having ingested 500 g of the rat bait Neosorexa or about 25 mg of difenacoum. Upon admission, she had a prolonged prothrombin time. She was treated with vitamin K for 45 days.

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The clotting activity returned to normal 30 days after the beginning of treatment (Barlow et al., 1982). Two cases of occupational exposure to difenacoum (and brodifacoum) have been reported by Park et al. (1986). Both patients were chronically exposed during manufacture of coumarin anticoagulants (brodifacoum, difenacoum, and warfarin). However, both difenacoum and brodifacoum were detected on several occasions in the first patient at plasma levels ranging from 30 to 50 ng/ml. They were not detected in the second patient 3 years after the last exposure, at which time he favorably responded to the vitamin K and fresh frozen plasma and was no longer professionally exposed. The metabolism of vitamin K was studied in both patients and in two female siblings of the first patient. In both patients, the plasma concentrations of vitamin K and vitamin K-2,3-epoxide after IV administration of 10 mg of vitamin K were measured. In the first patient, increased concentrations of the epoxide (300–500 ng/ml) were measured 2–24 h after vitamin K administration. This investigation was repeated several times over a period of 18 months and the epoxide concentrations were always found to be elevated, but the prothrombin time was always within the normal range. In the second patient, after his prothrombin time had been normal for 3 years, the vitamin K-2,3 epoxide was elevated on the two occasions where it was measured (100–200 ng/ml) after intravenous injection of 10 mg of vitamin K, whereas the clotting factors were in the normal range. Having eliminated a genetically determined disorder on the basis of the absence of detection of vitamin K-2,3 epoxide in the plasma of the first patient’s parents and two sisters, the authors discussed the other possible mechanisms of action without coming to a conclusion. Nighoghossian et al. (1990) have reported the case of 59-year-old patient accidentally exposed to difenacoum with an unusual expression of the coagulopathy. The patient developed an episode of subacute tetraparesis following a severe and sudden neck pain, which the clinical examination showed to be resulting from a subdural cervical hematoma. Prothrombin complex was low and difenacoum was detected in the plasma. Recent reports of cases of poisoning indicate that selfpoisoning was the most frequent cause of intoxication (McCarthy et al., 1997; Soubiron et al., 2000; Terneu, et al., 2003). The coumarin anticoagulant difenacoum was detected by HPLC with multiwavelength ultraviolet detection in plasma from a 41-year-old man who presented with a severe deficiency of vitamin K-dependent clotting factors of unknown etiology. Plasma concentrations of difenacoum declined from 0.97 to 0.11 mg/l in 47 days with a terminal half-life of 11.7 days. Subsequently, plasma concentrations of difenacoum and descarboxyprothrombin unexpectedly increased. Seven months after exposure, clotting times were still prolonged. The patient continued to have episodes of epistaxis, hematoma, purpura, and bruising and he required

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frequent treatment with fresh-frozen plasma in addition to oral phylloquinone (200 mg/day). Intermittent and unexpected increases in plasma concentrations of difenacoum and descarboxyprothrombin suggested that covert, repeated ingestion of the anticoagulant was the most likely cause of the poisoning. The measurement of low concentrations of plasma phylloquinone except following supervised ingestion of the vitamin indicated that as an outpatient, the subject was not compliant with treatment despite his protestations to the contrary. He continued to deny this even when confronted by laboratory findings and at no time did he ever admit to self-poisoning (McCarthy et al., 1997). (a) Treatment of Poisoning In case of threatening hemorrhage, transfusion of fresh blood or fresh-frozen plasma is the initial step. Intravenous and oral administration of vitamin K for a prolonged period of time (several weeks) with regular monitoring of coagulation is necessary.

100.5.9  Bromadiolone 100.5.9.1  Identity, Properties, and Uses (a) Chemical Name 3-[3-Bromo[1,1’-biphenyl-4-yl)-3-hydroxy-1-phenylpropyl]-4-hydroxy-2H-1-benzopyran-2-one is the chemical name. The CAS registry number is 28772-56-7. (b) Structure See Figure 100.3. (c) Synonyms Bromadiolone is the approved common name (BSI, E-ISO, F-ISO). Trade names include Deadline, Lanirat, Maki, and SuperCaid. (d) Physical and Chemical Properties Bromadiolone has the empirical formula C30H23BrO4 and a molecular weight of 527.4. The technical material is a yellowish powdered mixture of two diastereoisomers of a minimum purity of 97% and a melting point of 200–210°C. Its water solubility is of the order of 20 mg/l at 20°C. Solubility in organic solvents is (20°C) 730 g/l for dimethyl formamide, 25 mg/l for ethyl acetate, 8 mg/l for ethanol. It is stable under normal storage conditions. (e) History, Formulations, and Uses Bromadiolone is a second-generation anticoagulant of the hydroxy-4-coumarin that was patented in 1967 for the control of commensal rats and mice, including those resistant to warfarin and first-generation anticoagulants, voles, and water voles. It is formulated as ready-to-use cereal and paraffin-based baits containing 0.005% bromadiolone.

Hayes’ Handbook of Pesticide Toxicology

100.5.9.2  Toxicity to Laboratory Animals Single-dose acute toxicity of bromadiolone is of the same order of magnitude as other second-generation anticoagulants: in the wild rat (Rattus norvegicus) 1.1–1.8 mg/kg and in the mouse (Mus musculus) 1.75 mg/kg (Buckle et al., 1972). The toxicity to nonrodent species has been reported to be 0.3 mg/kg in the rabbit, 25 mg/kg in the cat, 0.15–1.0 in the dog, and 0.5–2.0 in the pig (Meehan, 1984). Birds appear to be somewhat less susceptible, with acute oral LD50s of 1600 mg/kg for quails (Tomlin, 1994). The acute percutaneous LD50 is reported to be 2.1 mg/kg in the rabbit (Tomlin, 1994). The kinetic behavior of bromadiolone was investigated in the rat by Kamil (1987). Groups of four rats (Rattus norvegicus) were orally administered bromadiolone at single doses of 0.8 or 3 mg/kg bw and then serially sacrificed up to 97 h after dosing. Bromadiolone was analyzed by HPLC in the plasma, liver, and kidney. The elimination half-life was calculated at 25.7 h for the 0.8-mg/kg bw dose and at 57.5 h for the 3-mg/kg bw dose. Concentration in the liver was rapidly established and was found to be 14–46 times higher than those in the plasma. A series of toxicological, residue, secondary hazard, and environmental fate studies on bromadiolone were conducted by Poché (1988). The results of these studies showed that bromadiolone is rapidly eliminated from orally treated rats (Rattus norvegicus) and mice (Mus domesticus ), with 75% of the dose being eliminated within 4 days in the rat. Oral LD50 in beagle dogs is 10.7 mg/kg bw in males and 6.3 mg/kg bw in females. Concentrations of bromadiolone in carcasses of rodents collected in the fields after a baiting campaign were 1.92 ppm in Rattus rattus, 1.17 ppm in Mus domesticus, and 0.49 ppm in Spermophilus beecheyi. The results of field tests with grain and pelleted baits showed that bromadiolone degraded by 78 and 45% respectively over a 21-day period. Primary and secondary toxicity of bromadiolone was investigated in laboratory experiments on earthworms (Eisenia fetida) exposed to bromadiolone containing granulated baits for 14 days (Kriskova et al., 2007). A group of common voles (Microtus arvalis) was fed the same bait granules for 5 days to evaluate the level of primary toxicity. Subsequently, the earthworms were fed to a second group of voles also for 5 days to evaluate the secondary toxicity. Voles are predominantly herbivorous but in this experiment they were trained for 7 days prior to the beginning of the experiment to consume earthworms for practical reasons and laboratory convenience. No mortality was observed in either the earthworms or the voles. The bromadiolone concentrations in the livers of voles were comparable in both the “primary” and “secondary” groups: 2.34  0.10 g/g and 2.20  0.53 g/g, respectively, indicating, according to the authors, that the risk of secondary poisoning was not greater than the risk from primary poisoning from feeding on baits.

Chapter | 100  Rodenticides

The authors have also described the analytical method used to determined bromadiolone in the animal tissues. It is an HPLC method with electrochemical detection (HPLC-ED). (a) Toxicity to Wildlife A recent survey of the effects of anticoagulant rodenticides on nontarget wild species in France has shown that this category of compounds is responsible for a very limited number of identified causes of death in most species: 1–3% (Berny et al., 1997). Kupper et al. (2006) reported bromadiolone poisoning in foxes occurring in Northern Switzerland in the context of an inappropriate application of bromadiolone baits to control water voles (Arvicola terrestris). At least 40 foxes (Vulpes vulpes) were found dead. The cause of death and the nature of the compound were confirmed by the detection of bromadiolone both in the blood and in the thoracic and abdominal fluids. Several cases of wildlife poisoning with anticoagulant rodenticides, including bromadiolone, were reported in the state of New York (Stone et al., 1999). From 1971 through 1997, 51 cases were studied (49 occurred in the last 8 years of this period). Brodifacoum was implicated in 80% of the cases, diphacinone was detected in four cases, bromadiolone in three cases (once in combination with brodifacoum). Chlorophacinone and coumatetralyl accounted for three cases once each in combination with brodifacoum. Warfarin accounted for three cases occurring before 1989. One-half of the cases concerned secondary poisoning of predators, essentially owls, hawks, squirrels, raccoons, and deer. Poisoning of pets (dogs and cats) has also been reported involving bromadiolone, brodifacoum, and diphacinone (DuVall et al., 1989). (b) Mode of Action Bromadiolone is an anticoagulant with the same mechanism of action as the other second-generation rodenticides.

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simultaneous determination and the quantitation of 15 anticoagulants in human serum (brodifacoum, bromadiolone, difenacoum, difethialone, flocoumafen, etc.). The case of a 40-year-old woman who intentionally ingested four 42.5-g bags of 0.005% bromadiolone baits (equivalent to 8.5 mg of active substance or 0.17 mg/kg bw) 4 days prior to admission to the hospital was reported. No bleeding was observed upon admission but her prothrombin time, INR, and activated thromboplastin time (APT) were elevated. The first plasma bromadiolone level, 5 days after ingestion, was 92 ng/ml. Serial measurements of plasma bromadiolone levels demonstrated that bromadiolone follows elimination kinetics of a two-department model with a rapid decline phase (half-life, 3.5 days) followed by a slower termination phase (half-life, 24 days). A consistently normal coagulation picture was associated with a plasma level of bromadiolone less than 10 ng/ml without vitamin K treatment. Different values for the bromadiolone half-lives have been reported. The half-life in whole blood was estimated to be about 6 days in the initial phase of elimination by Vindenes et al. (2008) and about 10–13 days in the terminal phase. In this study, the mean plasma/blood ratio of bromadiolone was 1.7  0.6, indicating that in the case of poisoning, it is better to search for the responsible substance in plasma rather than in whole blood. Interestingly, the authors showed that after two freeze-and-thaw cycles of whole blood samples, the bromadiolone concentrations decreased. In this report, the analytical work was performed using a liquid chromatography-mass spectrometry (LC-MS) developed by the authors in whole blood. The limit of detection (LoD) is 0.005 mg/l and the limit of quantification (LoQ) 0.01 mg/l. (a) Treatment of Poisoning Intoxication by bromadiolone, like with other second-generation anticoagulant rodenticides, requires massive and prolonged administration of vitamin K.

100.5.10  Difethialone

100.5.9.3  Toxicity to Humans

100.5.10.1  Identity, Properties, and Uses

Human poisoning with bromadiolone seems to be uncommon with only occasional cases reported in the literature. The case of intentional ingestion of bromadiolone by a 55-year-old man hospitalized with a bleeding wound on his tongue was reported by Grosbosch et al. (2006). The International Normalized Ratio (INR) was elevated and bromadiolone was detected in his serum by liquid chromatography-electrospray ionization-mass spectrometry (LC-ESI-MS). The maximum serum concentration was 440 g/ml. Further sample analysis provided the calculation of an elimination half-life of 140 h. According to the authors, the analytical method used is suitable for the

(a) Chemical Name 3-[3-(4-Bromo[1,1’-biphenyl]-4-yl)-1,2,3,4-tetrahydro1-naphthalenyl]-4-hydroxy-2H-1-[benzothiopyran-2one] is the chemical name. The CAS registry number is 104653-34-1. (b) Structure See Figure 100.3. (c) Synonyms Difethialone is the approved common name (BSI). Trade names include Frap, Baraki, and Operats Plus.

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(d) Physical and Chemical Properties Difethialone has the empirical formula C31H23BrO2S and a molecular weight of 539.5. It forms a whitish/yellowish powder. It has a low vapor pressure (1.33  105 Pa at 22.6°C). It is practically insoluble in water: 0.39 mg/l at 25°C. Its solubility in ethanol, methanol, hexane, chloroform, and acetone in mg/l at 25°C is 0.7, 0.47, 0.2, 40.8, and 4.3, respectively. The octanol/water partition coefficient is 1.41  105. It is stable at temperatures up to 230°C. Its melting point is 236°C. It is highly sensitive to photolysis in aqueous solutions. It is strongly adsorbed in soils. (e) History, Formulations, and Uses Difethialone was introduced as a second-generation anticoagulant rodenticide in 1986 for the control of commensal rats and mice, including those resistant to first-generation anticoagulants. It is formulated as ready-to-use whole grain cereals and husked oat grain baits containing 0.0025% active substance.

100.5.10.2  Toxicity to Laboratory Animals The acute single-dose oral LD50 of technical grade difethialone (97.6% purity) has been reported as 0.56 mg/kg for rats, 1.29 mg/kg for mice, 4 mg/kg for dogs, and 2–3 mg/kg for pigs. The percutaneous acute LD50 was determined in rabbits at 5.3 mg/kg, 6.5 mg/kg in male rats, and 5.3 mg/kg for females. By inhalation, in rats exposed for 4 h, the LC50 is between 5 and 19.3 mg/m3. It is not irritant to rabbit skin and only slightly irritant to the eyes. It is not a skin sensitizer. It has no mutagenic or teratogenic potential. The only effects seen in a 90-day feeding study in the rat were those expected on blood coagulation. Comparative toxicity to nontarget species was investigated. Difethialone is less toxic to pigs than brodifacoum and warfarin: no mortality occurred 30 days after administration of 1 mg/pig/day, whereas 1 mg of brodifacoum/pig/day induced mortality after 14, 15, and 18 days and 5 mg warfarin/pig/day induced mortality after 8, 9, and 10 days of administration (Lechevin and Poche, 1988). In another comparative laboratory study, difethialone and warfarin were administered to gerbils (Meriones hurrianae Jerdon) and to Swiss albino mice. It was observed that difethialone had the shortest feeding period (1 day) and the lowest lethal dose (1.18 mg/100 g bw) for mice and 1.42 mg/kg bw for gerbils to obtain 100% mortality compared to warfarin; 5 days of feeding were necessary and the lethal dose was also higher (38.95 mg/100 g bw in mice and 151.82 mg/kg bw in gerbils) (Saxena and Sahni, 2001). Saravan and Kanakasabai (2004) have experimentally shown that difethialone had the potential to cause secondary poisoning in wildlife species. They fed difethialone poisoned rats to groups of captive barn owls (Tyto alba) for either 1, 3, or 6 consecutive days. Owls exposed for 1 and 3 days survived but in the group fed for 6 days, all birds died.

(a) Absorption, Distribution, Metabolism, and Excretion In rats treated by oral administration, difethialone has a short half-life in blood and a longer half-life in the liver. It is essentially eliminated in the feces as the unchanged parent compound, indicating a very limited metabolism (Tomlin, 1994). In the rat, difethialone has a short half-life in plasma and a longer one in the liver: 2.3 days and 108 days, respectively, following oral administration of a 0.5-mg/kg bw dose; at 5 mg/kg bw, the plasma half-life is 2.8 days and cannot be calculated in the liver because of mortality occurring 6 days after administration. Elimination is essentially fecal with almost no degradation. The fecal elimination was 83.1% in 4 days with a dose of 5 mg/kg bw (Lechevin and Poche, 1988). (b) Mode of Action Difethialone is an anticoagulant with the same mode of action as the other second-generation compounds.

100.5.10.4  Toxicity to Humans It seems that so far very few cases of human poisoning by difethialone have been reported in the literature, which does not automatically mean that the number of difethialone cases of poisoning is low. The case of poisoning of a 22-year-old male was reported by Vasquez and Rodriguez (2000). The man had repeatedly handled a difethialone-containing rodenticide in his house for several months before being admitted to the hospital with left lumbar pain and hematuria, epistaxis, and bleeding gums. Upon admission, the coagulation parameters were severely disturbed. The condition required a prolonged treatment with vitamin K and fresh frozen plasma. (a) Treatment of Poisoning Intoxication by difethialone, like with other second-generation anticoagulant rodenticides, requires massive and prolonged administration of vitamin K.

100.6  Vitamin D-related compounds 100.6.1  Ergocalciferol 100.6.1.1  Identity, Properties, and Uses (a) Chemical Name 9,10-Secoergosta-5,7,10(19),22-tetraen-3–01 is the chemical name. (b) Structure See Figure 100.5. (c) Synonyms The common name calciferol is approved by the BPC (British Pharmacopeia Commission); ergocalciferol is

Chapter | 100  Rodenticides

approved by USP (U.S. Pharmacopeia). It is also known as vitamin D2, but for safety reasons it is prohibited in some countries to mention the identity of these compounds as vitamin D on rodenticide labels [Food and Agriculture Organization (FAO), 1979]. Other names include activated ergosterol, Derat Concentrate, Deratol, and Hi-Deratol. The trade name is Sorexa C.R. for a combination of calciferol and warfarin. (d) Physical and Chemical Properties The empirical formula is C28H44O and the molecular weight is 396.63. It forms colorless prismatic crystals. The melting point is 115–118°C. It is insoluble in water and soluble in most organic solvents; solubility at 7°C is 69.5 g/l in acetone. It is slightly soluble in vegetable oils. Deterioration of pure, crystalline vitamin D2 is negligible after storage for 9 months in evacuated amber ampules at refrigerator temperature. However, calciferol tends to decompose in the presence of air and moisture. The stability of corn oil solutions is, however, satisfactory (Greaves et al., 1974). (e) History, Formulations, and Uses The rodenticidal properties of calciferol were described by Greaves et al. (1974). The commercial rodenticide was introduced in the United Kingdom in 1974 as a combination of calciferol and warfarin formulated as a ready-touse bait on canary seed (1 g of calciferol and 250 mg of warfarin per kilogram). It is also available as an oil concentrate (20 g/l). Calciferol is used as a rodenticide for control of commensal rats and mice. Toxicity tests with calciferol combined with warfarin suggest that an additive effect between the compounds could exist. One interesting advantage of calciferol is that it is toxic to warfarinresistant rodents. A second advantage is that it kills more rapidly – within 1 week instead of the 1–3 weeks that are often required with anticoagulants (Greaves et al., 1974).

100.6.1.2  Toxicity to Laboratory Animals (a) Basic Findings The acute oral LD50 of calciferol is 56 mg/kg for rats and 23.7 mg/kg for mice. When administered daily for 5 consecutive days to laboratory rats, the LD50 falls to 7 mg/kg/day. Lethal doses have also been reported by Gill and Redfern (1979) for the multimammate rats Mastomys natalensis; they range from 78 to 107 mg/kg (mean, 96 mg/kg) in males and from 108 to 137 mg/kg (mean, 119 mg/kg) in females when administered for 1 day at a concentration of 0.1% in bait. Similar values were reported by Greaves et al. (1974) for wild rodents. When calciferol was given at doses of 100 mg/kg by stomach tube on 1 or more days, laboratory rats and mice became visibly ill within 3 days. The clinical signs are characterized by loss of appetite, listlessness, piloerection, hunched position, absence of

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reaction to external stimuli, weight loss, priapism, and frequent micturition (Gillman et al., 1960). (a) Absorption, Distribution, Metabolism, and Excretion The action of calciferol is to raise blood calcium levels by stimulating the absorption of calcium from the intestine and mobilizing skeletal reserves. This mechanism is slow and takes many hours to build up an effective level; the period of latency between the ingestion of calciferol and the development of hypercalcemia and occurrence of lethal lesions is of the order of several days, usually 4 or 5 (Greaves et al., 1974). Calciferol has a long biological half-life in mammals and hypercalcemia induced by overdosage may continue for 6–9 months (Buckle et al., 1972). The various tissues of the body store vitamin D for varying periods. The depletion of stores in the body is caused partly by its fecal excretion and partly by its destruction in the body. (c) Biochemical Effects Calciferol is the most important factor for the optimal absorption of calcium. It is responsible for the synthesis of a protein that binds calcium in the intestinal mucosa, especially in the duodenum. In the absence of calciferol, calcium cannot be absorbed. (d) Effects on Organs and Tissues All forms of vitamin D are toxic when given in sufficiently large amounts. Excessive doses of calciferol mobilize the phosphorus and calcium from the tissues, thus broadly having an opposite effect to normal doses. The soft tissues tend to become calcified while the bone tends to be rarefied. The soft tissues most affected are the renal tubules and the media of the small renal arteries and of the large vessels, especially the aorta. The bronchi, lungs, heart and coronaries, and the stomach also are affected. In dogs, there is atrophy of the testes and the prostate while the parathyroids are smaller than normal. The histochemical effects of vascular injuries induced by calciferol orally administered to male and female Wistar rats for 5 consecutive days at doses ranging from 25,000 to 150,000 IU (1 mg of calciferol is equivalent to 40,000 IU) were studied by Gillman et al. (1960). By day 15 of the experiment – that is, day 10 after the last day of calciferol administration – necrosis of the spleen was observed in many rats that died or were sacrificed. Among the survivors, there were indications that the damaged spleen had been regenerated completely by day 25. The heart was severely injured very early in the experiments. Similarly, the coronary arteries were observed showing early dilation, injury to the internal elastic membranes, and associated early calcification. These changes were associated with sclerotic repair. In the aorta, there

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was an apparent relationship between the intensity of the reaction and the doses of calciferol received by the rats. By day 20, a very large concentration of calcium accumulated in the aorta (13–14% calcium compared to 22% in the femur). Similar observations were made by Grant et al. (1963), who also showed that this phenomenon was a three-step process involving early widespread alterations in many organs and tissues, followed by spontaneous recurrent resolution and reappearance of calcification for many months, even after a single episode of acute intoxication. These authors also showed distinct differences in the time of onset and rate and extent of calcification of various tissues. Nikodemusz et al. (1981) have shown that in six male and six female common voles (Microtus arvalis), single acute oral dosing with calciferol (80–620 mg/kg) induced morphological changes representing varying degrees of parenchymal degeneration and calcification in the kidneys, lungs, and heart. Calcium deposits were observed in the esophageal and gastric mucosae, as well as in the aortic media. These lesions were not observed in the group receiving 80 mg/kg of calciferol. The approximate lethal dose by the oral route was estimated as 120 mg/kg in males and 280 mg/kg in females. The survival time ranged from 53 to 94 h. Tarrant and Westlake (1984) have also shown that feeding laboratory-reared male rats (Rattus norvegicus) and male and female quail (Coturnix coturnix japonica) a diet containing 0.1% calciferol (the recommended field concentration) for 2 days induced in both species a similar pattern of calcium deposits in the kidneys, beginning on the 2nd day after the animals were returned to a standard, calciferol-free diet. (e) Effects on Reproduction Adult female New Zealand White rabbits had been intramuscularly treated with ergosterol in cottonseed oil in divided doses every other day for a total of 1.5 million IU. Three other groups of five rabbits each received intramuscular injections of ergosterol daily for the duration of the gestational period for a total of 2.5–3.5 million and 4.5 million IU. At autopsy all the females given 2.5 million IU and above died spontaneously within 65 days after their first injection of calciferol and all that were pregnant aborted during the first 12 days of pregnancy or delivered macerated fetuses. All aortas had various degrees of pathological changes, including those from females treated with 1.5 million IU. A total of 14 abnormalities of the aorta were observed in the 34 offspring whose mothers had been treated with excessive levels of ergocalciferol. Aortic lesions that appeared similar to the supravalvular aortic stenosis seen in humans were noted in six rabbits. The blood levels of ergocalciferol in the mothers and their offspring were seven and nine times greater than those in the control classes and their offspring, respectively, indicating that transplacental passage occurred (Friedman and Roberts, 1966). A similar experiment was carried out by

Hayes’ Handbook of Pesticide Toxicology

Friedman and Mills (1969) on 15 pregnant New Zealand White rabbits given divided doses of intramuscular ergocalciferol every other day, starting on the 2nd day of insemination and throughout the pregnancy for a total of 750,000 IU. Characteristic malformations were observed in the offspring. These were represented by premature closure of cranial structures, small skulls compared with the controls, and narrowing of the body of the mandible. The maxillary and mandibular central incisors showed severe enamel hypoplasia. Some cases of anodontia and abnormal palatal shape were noted. Similar craniofacial malformations are frequently associated with supravalvular aortic stenosis in children. Several additional studies of the effects of vitamin D on reproduction and fetal development in rats and rabbits have been reported. Latore (1961) found that excess ergocalciferol reduces fertility in rats when administered on days 0–7 of gestation but not when administered on days 8–21. Nebel and Ornstein (1966) have demonstrated that ergocalciferol administered to rats (Rattus rattus) in daily doses of 20,000 IU for 1–3 weeks significantly affected the genital cycle, fertility, and early pregnancy from both the morphological and functional points of view in direct relation to the beginning and duration of ergocalciferol administration. Ornoy et al. (1968) showed that vitamin D2 (ergocalciferol) given to pregnant albino rats at daily doses of 4000, 20,000, or 40,000 IU from day 9 to day 21 of pregnancy crosses the placental barrier and induces alterations of the mineral composition in the fetal bones at the 40,000 IU level only. In the same treated group, placentas, fetuses, and fetal bones were found to be smaller than those in the untreated groups. However, it seems from the experimental data that pregnant rats are more tolerant to rather high levels of ergocalciferol than nonpregnant females (Ornoy et al., 1968; Potvliege, 1962). (f) Factors Influencing Toxicity In commercially available preparations, calciferol is often associated with warfarin in a 4:1 ratio because it has been shown that this mixture produced a marked increase in mortality in Norway rats (Greaves et al., 1974), thus suggesting an additive effect of the two compounds. At the LD50 level, toxicity of the mixture is intermediate between the toxicities of calciferol and warfarin. (g) Treatment of Poisoning in Animals It seems that the only case of poisoning of a domestic animal reported is that of a 4-year-old boxer dog seen several days following ingestion of a warfarin-calciferol mixture. Despite treatment with vitamin K, antibiotics, and vitamins, the end of the tongue became necrotic and had to be removed and large areas of the skin were also necrosed, indicating generalized vascular damage. The dog made a slow recovery (Edlin, 1982). It is likely that treatment with vitamin K had prevented the hemorrhagic manifestations of warfarin toxi­city from showing, but that the vascular injuries due to calciferol were responsible for the necrotic effects.

Chapter | 100  Rodenticides

100.6.1.3  Toxicity to Humans (a) Therapeutic Use As a vitamin, calciferol is used to prevent and to cure rickets, tetany, spasmophilia, and osteoporomalacia. Several decades ago, vitamin D was recommended for the treatment of lupus vulgaris (skin tuberculosis limited to the face); massive doses were used, sufficient to provoke a mild degree of hypervitaminosis. The purpose was to induce calcification of the subcutaneous lesions in order to stop their progression. In the case of rickets, the recommended preventive dose was 500–1500 IU/day, and the curative dose was 1000–3000 IU/day; it was estimated that 10,000 IU represents a toxic dose. The minimal toxic overdose does not appear to be many times greater than the optimum curative dose (Harris, 1955). Many cases of vitamin D2 as well as D3 poisoning in humans have been reported. The oldest ones were analyzed by Bicknell and Prescott (1953). More recent cases have been reported, all related to accidental or inadvertent overdosing. A 69-year-old woman with hypoparathyroidism was treated with a twice-weekly dose of 50,000 IU of calciferol. However, for 4 weeks before admission to hospital she had mistakenly taken 300,000 IU (7.5 mg) a day. On examination she was lethargic and mentally confused, with muscular hypotonia. In this report, two other similar cases concerning elderly women are reported; one had received 100,000 IU (2.5 mg) calciferol orally on alternate days for 3 months and the other had received 400,000 IU (10 mg) a day by mouth and 600,000 IU of calciferol once a week by injection during the 2 months before admission to hospital. All three patients had elevated serum calcium. All three of them were treated with intravenous injections of porcine calcitonin, which caused serum calcium to fall back to a normal level within 2–3 days (Buckle et al., 1972). Davies and Adams (1978) have also reported eight cases of severe vitamin D poisoning. In six patients, the therapy was unnecessary and in two others inadequate supervision of treatment resulted in overdosage. Paterson (1980) reported 21 cases of hypercalcemia due to vitamin D poisoning; among them two patients died while intoxicated. Overall, the clinical symptoms induced by chronic poisoning include, in various associations and intensity, loss of appetite, loss of weight greater than would be expected from the loss of appetite, nausea, vomiting, and constipation or diarrhea. Abdominal pain may be so severe as to lead to unnecessary laparotomies. Headaches are usual and one special form has been noticed. This is a tightness across the back of the head which progresses to acute sensitiveness of the scalp. Mental confusion and loss of memory may also be seen. Epileptiform fits are a rare complication. Metastatic calcification has been described along with vascular and renal calcification. An unusual, albeit typical case has been reported by Cohen et al. (1979): a case of deafness due to a long-term overdosage for the treatment of pseudohypoparathyroidism (2.5 mg of calciferol daily

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for 4 years). The patient had a 3-month history of deafness, weight loss, anorexia, and weakness. She had extensive calcification of the tympanic membranes, corneas, kidneys, and blood vessels. She had also, of course, high serum calcium, which was successfully treated with calcitonin and prednisolone and a low-calcium diet. She was discharged from hospital symptom-free except for the deafness. Hypercalcemia is the earliest sign of vitamin D overdosage. It was suggested at one point that increased intake of vitamin D through consumption of fish liver in Norway might be correlated with an increased probability of myocardial infarction (Linden, 1974). However, this hypothesis, based on a retrospective study of a number of patients with myocardial infarction, angina pectoris, and degenerative joint diseases, was later criticized because of serious bias and shortcomings in the methodology (Lindahl and Lindwall, 1975). A prospective study, conducted in the same area of northern Norway, has not confirmed the existence of a higher risk of myocardial infraction related to vitamin D intake or status of the population (Vik et al., 1979). (b) Use Experience and Dosage Response There is no known report of accidents occurring with calciferol used as a rodenticide. It is difficult to estimate the minimum toxic dose in humans; however, from the reported cases of overdosage, it would seem that 0.15 mg/kg/day for 3 weeks may lead to clinical signs of poisoning. (c) Treatment of Poisoning Treatment of acute and chronic overdosages with calciferol requires that hypercalcemia be brought down to a normal level quickly; intravenous injection of calcitonin is the specific treatment to be applied under close monitoring of the serum calcium level. Steroid therapy is also often effective but slower, normocalcemia being achieved after 5–7 days. Other methods are available but they are not devoid of problems: chelation with sodium ededate is only temporary in its effect and is nephrotoxic. Sodium phosphate is also effective but may be responsible for metastatic calcifications (Buckle et al., 1972). One case of vitamin D2 poisoning in an elderly woman was successfully treated by inducing the hepatic microsomal enzymes with 500 mg/day of glutethimide. The mechanism of action is still unknown (Iqbal and Taylor, 1982), but this approach is rather slow since the serum calcium level did not fall back to normal values until day 12 of treatment.

100.6.2  Cholecalciferol 100.6.2.1  Identity, Properties, and Uses (a) Chemical Name 9,10-Secocholesta-5,7,10(19)-trien-3-betaol is the chemical name. (b) Structure See Figure 100.5.

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CH3 CH3 CH3

CH3

CH3

CH3

CH3

CH2

CH3 CH3

CH2 HO

HO Ergocalciferol

Cholecalciferol

Figure 100.5����� ���� Two forms of vitamin D used as rodenticides.

(c) Synonyms Some synonyms are activated 7-dehydrocholesterol, oleovitamin D3, cholecalcifero, natural vitamin D3. The trade name for the rodenticide is Quintox. (d) Physical and Chemical Properties The empirical formula is C27H44O and the molecular weight is 384.62. It forms tine needles. The melting point is 84–86°C. It is practically insoluble in water, soluble in the usual organic solvents, and only slightly soluble in vegetable oils. It is oxidized and inactivated by moist air within a few days. However, the formulated product is stable over 1 year at ambient temperatures in sealed packages. (e) History, Formulations, and Uses Cholecalciferol is formulated for use as a rodenticide as a grain bait containing 750 ppm (0.075%) of active ingredient and commercialized under the trade name Quintox. It is a single-feeding and multifeeding rodenticide used for controlling anticoagulant-resistant rats and mice. There is a period of time between feeding and death similar to but perhaps shorter than that observed with anticoagulant rodenticides. The rodenticidal activity of cholecalciferol is comparable to that of ergocalciferol.

100.6.2.2  Toxicity to Laboratory Animals The two forms of calciferol are equally toxic to most mammals. The acute oral LD50 of cholecalciferol is 43.6 mg/kg for Rattus norvegicus and 42.5 mg/kg for mice (Mus musculus). In the dog, the oral LD50 is 88 mg/kg. Two groups of two dogs each were administered amounts of cholecalciferol-containing baits equivalent to 20 and 10 mg cholecalciferol/kg bw or approximately onefourth and one-eighth of the published LD50, respectively. All dogs developed hypercalcemia and hypophosphatemia and died. Major lesions were gastrointestinal hemorrhage, myocardial infarction, and mineralization of vascular walls

(Gunther et al., 1988). These values of 10–20 mg/kg were confirmed by El Bahri (1990). Therefore, the acute LD50 of cholecalciferol in dogs is lower than previously believed. (a) Absorption, Distribution, and Excretion Cholecalciferol after absorption from the intestine is transported to the liver, where it is metabolized to 25-hydroxycholecalciferol by an NADPH-dependent reaction. This metabolite is then transferred to the kidney and converted to 24-, 25-, or 1,25-dihydrocholecalciferol by mitochondrial mixed-function oxidases (McClain et al., 1980). After their intestinal absorption, ergocalciferol (vitamin D2) and cholecalciferol (vitamin D3) undergo an identical metabolic C pathway (Fournier et al., 1985). The metabolism and pharmacokinetics of one metabolite of cholecalciferol, 24,25-dihydrocholecalciferol have been reviewed by Jarnargin et al. (1985); the excretion curve shows an initial fast phase with a plasma half-life of 0.55 h and a second slow phase with a plasma half-life of 73.8 h in the rat. The clearance from plasma, liver, and kidney but not intestine follows a two-compartment model. The most potent form of vitamin D3, 1,25-dihydrocholecalciferol, has been shown to be responsible for the stimulation of intestinal absorption of calcium and the metabolism of calcium in bone. However, Frolick and Deluca (1973) have shown that when given orally to rats, 1,25-dihydrocholecalciferol is rapidly modified during its passage through the intestine, thus reducing its physiological activity to a large extent. Orally administered high doses of cholecalciferol (37.5 g/day for 14 days followed by an 88-day observation period) to female Wistar rats rapidly increase the plasma cholecalciferol level to reach a steady-state. Plasma 25-hydroxyvitamin D (25-OH-D) and adipose tissue levels of cholecalciferol increased linearly for 2 days after treatment. Subsequently, half-lives of plasma cholecalciferol and 25-OH-D and perirenal and subcutaneous adipose tissue were: 1.4, 22.5, 97.5, and 80.9 days, respectively. Fasting, as compared with ad libitum feeding, caused

Chapter | 100  Rodenticides

increased plasma free fatty acids, weight loss and increased adipose tissue cholecalciferol. It did not affect plasma cholecalciferol immediately after treatment but raised 25-OH-D. Fasting at the end of the study decreased plasma cholecalciferol and increased 25-OH-D. The authors concluded that orally administered cholecalciferol rapidly accumulates in adipose tissue and that it is very slowly released (Brouwer et al., 1998). (b) Effects on Organs and Tissues Moderately excessive doses of cholecalciferol, 100, 300, 2000, and 4000 IU/kg of feed, given for 4 months to experimental Yorkshire pigs rapidly produced gross arterial lesions (fibromuscular interstitial thickening of the coronaries, especially at the branching sites). Macrophages, plasma cells, and mast cells were observed to accumulate in the subendothelial space. The extent of these lesions, resembling those commonly seen in atherosclerosis in humans, was more or less dose-related (Toda et al., 1985). Since both forms of vitamin D (ergocalciferol and cholecalciferol) follow the same metabolic pathway, it is expected that they are responsible for inducing the same lesions. (c) Effects on Reproduction Calcitriol (1,25-dihydrocholecalciferol), which is the most biologically active metabolite of cholecalciferol, has been administered to pregnant rats and rabbits at daily doses of 0.02–0.08 and 0.30 g/kg from days 7–15 of gestation in the rat and from days 7–18 in the rabbit. In rats no adverse effect on fertility, litter parameters, and offspring was observed. Hypercalcemia and hypophosphatemia were observed in pregnant rats of the middle- and highdose groups, as well as hypercalcemia in pups. Calcitriol induced maternal mortality and fetotoxicity in rabbits treated with 0.3 g/kg/day. Two litters at the highest dose and one litter at the middle-dose level contained fetuses with multiple abnormalities (McClain et al., 1980).

100.6.2.3  Toxicity to Humans (a) Experimental Oral Exposure The active metabolite 1,25-dihydrocholecalciferol was administered either orally or intravenously to four healthy volunteers and three patients with hypoparathyroidism. After an oral dose, the highest serum concentration of radioactivity was reached after 4 h. The route of administration had little apparent effect on the serum concentration or on the rapid phase of elimination, but the slow phase of excretion was longer after oral administration. The highest urinary excretion rate was observed during the first 24 h and was little affected by the route of administration. The half-life of 1,25-dihydrocholecalciferol was 3–5 days. On average, 40% of the dose of 1,25-dihydrocholecalciferol was excreted within 10 days (Mawer et al., 1976).

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(b) Therapeutic Use Cholecalciferol being the natural form of vitamin D, the therapeutic uses are those outlined for ergocalciferol. (c) Accidents and Use Experience These are the same as those mentioned for ergocalciferol. There is no known report on a human accident specifically attributable to cholecalciferol used as a rodenticide. However, intoxication from overdosage during therapeutic use of vitamin D3 is well known and signs and symptoms have been well studied (Navarro et al., 1985).

100.6.2.4  Toxicity to Nontarget Species There are reports of accidental poisonings in nontarget species in the literature. It seems that dogs are essentially concerned. Fooshee and Forrester (1990) report an accidental poisoning of two dogs. One dog died shortly after admission and the second dog successfully responded to treatment with sodium chloride solution, prednisolone, furosemide, and calcitonin, but treatment was needed for a longer period than anticipated as the calcemia did not stabilize for about 1 month. Poisoning of a 27-kg female German shepherd dog was reported by Scheftel et al. (1991). An initial serum calcium concentration of 15.7 mg/l was successfully reduced to normal during a 10-day treatment with calcitonin and prednisolone. Diagnosis of cholecalciferol toxicity in a young dog was made on the basis of exposure history and the finding of elevated serum calcium levels. Fluid therapy, diuresis, and corticoids lowered the serum levels over a period of 5 days but during this period the dog showed clinical signs of severe depression, anorexia, and vomiting and serum calcium levels rose on day 6, remaining high up to day 9 at which point the animal had to be euthanized. Autopsy revealed diffuse metastatic mineralization throughout the body, particularly the lung, kidney, heart atria, and stomach. The amount of cholecalciferol ingested was determined to be well below the reported lethal dose in dogs, indicating that young dogs may be more susceptible than initially thought (Talcott et al., 1991). It should be noted that in the present case, calcitonin was not administered, which may be one reason for the deterioration of the dog’s condition. A cat suspected of having ingested a cholecalciferolcontaining rodenticide was seen with hypercalcemia and hyperkalemia and was acidotic. Despite management of hypercalcemia and preservation of renal function with physiologic saline solution, furosemide, dopamine, and calcitonin, the cat died, apparently as a result of extensive pulmonary mineralization (Peterson et al., 1991).

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100.7  Miscellaneous synthetic organic rodenticides As far as is known, the miscellaneous synthetic organic rodenticides (see Figure 100.6) are unrelated to one another or to other groups of pesticides either pharmacologically or chemically. The two that have been studied in humans are remarkable. Chloralose is an anesthetic, albeit one with the property of inducing myoclonic seizures. With the possible exception of this compound, anesthetics have nothing to offer for killing vertebrate pests. Far too many animals become anesthetized before consuming a fatal dose, and they later recover. In fact, the possibility of recovery seems to be taken into account in the British practice of recovering birds affected by the compound. Norbormide is of interest as one of the most selective poisons known. Its limitation is not any lack of toxicity to species of the genus Rattus but the problem of secondary bait refusal (Greaves, 1966).

(d) Physical and Chemical Properties Chloralose has the empirical formula C8H11Cl3O6 and a molecular weight of 309.54. It forms a crystalline powder that melts at 187°C. It is soluble in ether and glacial acetic acid, slightly soluble in chloroform, and almost insoluble in petroleum ether. Its solubility in water at 15°C is 0.44%. Chloralose reduces Fehling’s solution only after prolonged heating. It is hydrolyzed into its two components by acids. (e) History, Formulations, and Uses Chloralose has been in use in Europe for many years. Its narcotic properties are employed to immobilize depredating birds and render them easier to kill by other means. Baits contain about 1.5% of the active compound. It has been reported that chloralose is also used on seed grain as a bird repellent. It is used against mice in baits of up to 40 g/kg. All baits should contain a warning dye. Chloralose is used in medicine as a soporific and formerly was used as an anesthetic.

100.7.1  Chloralose

100.7.1.2  Toxicity to Laboratory Animals

100.7.1.1  Identity, Properties, and Uses

The oral LD50 values of chloralose in rats and mice are in the range of 300–400 mg/kg. Cats are more susceptible (100 mg/kg) and dogs more resistant (600–1000 mg/kg) (Cornwell, 1969). The compound is more toxic to many birds than to mammals. Oral LD50 values have been determined for the starling (75 mg/kg), redwing blackbird (32 mg/kg), yellow-headed blackbird (133 mg/kg), crow (42 mg/kg), pigeon (178 mg/kg), house finch (56 mg/kg), house sparrow (42 mg/kg), mallard duck (42 mg/kg), mourning dove (42 mg/kg), and white-crowned sparrow (56 mg/kg) (Schafer, 1972). Hanriot and Richet, who advocated chloralose as a soporific, found that dogs survived oral dosages as high as 610 mg/kg but were killed by dosages of 660 mg/kg and greater. Cats survived oral dosages of 65 mg/kg or lower but one was killed by a dosage of 71 mg/kg and all were killed by 140 mg/kg or more. Dogs survived intravenous injection of 120 mg/kg or less but were killed by 150 mg/kg. A dosage of 12.5 mg/kg produced severe symptoms in a cat (Hanriot and Richet, 1897). After some delay, animals poisoned by chloralose show incoordination, vertigo, tremor, and failure to recognize objects. The sense of pain is lost but there is increased reactivity to touch, sound, or electric shock. If stimulated, the animals respond reflexively and with full force. If artificial respiration is withheld, animals that have received a sufficient dosage die of respiratory failure (Hanriot and Richet, 1897). Thus it was recognized very early that although response to chloralose is somewhat similar to that to chloral, it is also similar to the response to strychnine in the sensitization to external stimuli. Chloralose is metabolized to chloral, CH(OH)2-CCI3 (Cornwell, 1969), oxidized to trichloroacetic acid, and

(a) Chemical Name 1,2-O-(2,2,2-trichloroethylidene)--D-glucofuranose is the chemical name. (b) Structure See Figure 100.6. (c) Synonyms Chloralose (BSI) is the common name in use. Other nonproprietary names include -chloralose, -d-glucochloralose, anhydroglucochloral, chloro-alosane, and glucochloral. Trade names include Alphakil and Somio. The CAS registry number is 15879-93-3.

N C

—

C—O

CH– CCl3

—

—— ——

—

H—C—O

O

—

HO —C —H —

H—C —OH —

H—C—OH

C N

—

H—C—OH

NH OH

H

α-Chloralose

Norbormide

Figure 100.6  Miscellaneous synthetic organic rodenticides.

O

Chapter | 100  Rodenticides

reduced to trichloroethanol (Marshall and Owens, 1954; Owens and Marshall, 1955). The latter metabolite is responsible for much of the hypnotic effect of chloral hydrate; all tissues studied so far are capable of forming it from chloral hydrate (Butler, 1949). Trichloroethanol combines with glucuronic acid in the liver to form the pharmacologically inactive urochloralic acid, which is readily excreted in the urine (Lees, 1972). Chloralose was not tumorigenic in two strains of mice that received it at the highest tolerated level for 18 months (Innes et al., 1969). (a) Treatment of Poisoning in Animals Administration of analeptic drugs or stimulants of the central nervous system such as methylamphetamine (0.5–4 mg/kg of body weight, orally or intramuscularly) or ephedrine (2.5 mg/kg of body weight subcutaneously) has been recommended and successfully applied in poisoned dogs (Bennett, 1972; Smith and Boyd, 1972), although this had already been criticized (Shepherd, 1971) on pharmacological and biochemical grounds. In addition, supportive therapy to correct any hypothermia and respiratory problems may be indicated in severely poisoned animals.

100.7.1.3  Toxicity to Humans (a) Therapeutic Use Chloralose was introduced in 1888 and 1893 as an anesthetic and soporific. Its anesthetic use was soon dropped, presumably because an effective dosage tended to cause excessive muscular activity. However, this same feature was regarded as an advantage in connection with a soporific. Thus Sollman (1901) considered chloralose preferable to chloral except for insomnia due to exaggerated reflex irritability. He pointed out that chloralose is a stronger hypnotic, heightens the reflexes, has less action on the heart, and produces practically no local irritation. Chloralose has gone out of fashion in the United States because its action is somewhat delayed compared to that of chloral (Sollman, 1942) and perhaps because activation of reflexes was not considered an appropriate property of a soporific. (b) Accidental and Intentional Poisoning Most reported cases of poisoning by chloralose have occurred in France, where the compound is used medically as a soporific and sedative and also is used as a poison to kill crows and rats. Tempé and Kurtz (1972) listed 60 brands of poison based on chloralose, of which 31 also contained ANTU (see Section 100.4.1.3). In their series of 22 acute intoxications by chloralose, only one was caused by the compound intended for use as a drug; the others were caused by rat poisons. Most cases of poisoning have involved attempted – generally unsuccessful – suicide. A few cases of mild

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accidental poisoning of children have been recorded (Gaultier et al., 1962). The characteristic effect of chloralose is coma, which may be preceded by vomiting, vertigo, trembling, and a sensation of inebriation. In massive intoxication, the coma may appear in some minutes but usually appears in one to several hours after ingestion of chloralose (FavarelGarrigues and Boget, 1968; Tempé and Kurtz, 1972). The patient may be calm and limp or may be agitated. Cornette and Franck (1970) saw only cases in agitated coma with varying degrees of myoclonia. The myoclonia was always reinforced by stimulation and it occurred predominantly in the arm or leg that was stimulated. There was a bilateral, symmetrical seizure, occasionally with hypersalivation and incontinence of urine. However, tonic spasm or tonic clonic convulsions were not seen. These authors never encountered a case of hypotonic or “calm” coma seen by others. They attributed the difference to better treatment, but it would seem difficult to exclude differences in dosage as a cause. Favarel-Garrigues and Boget (1968) specifically noted that this form of coma usually but not always occurred in massive intoxications, and one almost always saw hyperreactivity, clonic jerks, and a state of agitation in such cases during recovery. Tempé and Kurtz (1972) reported similar experience. The reflexes are active in agitated coma. In nine of the 22 cases reported by Tempé and Kurtz (1972), six showed a positive Babinski test bilaterally. Reflexes are diminished or absent in massive intoxication. Authors have not agreed on whether the most severe seizures caused by chloralose are truly epileptic (Moene et al., 1969) or are bilateral, synchronous myoclonic disturbances (Cornette and Franck, 1970). That the latter view is correct seems to depend not on any difference in clinical severity but on a lack of correlation of the EEG with the physical disturbance. Hypersecretion of the respiratory tract apparently is the most life-threatening aspect of intoxication. It may occur in the absence of ANTU but is more common and more severe when ANTU is involved. This hypersecretion may appear early, suddenly, and severely, but it also disappears in less than an hour. Unlike pulmonary edema, the secretion is not sanguineous and it is poor of albumin. The mild appearance on x-ray contrasts with the clinical severity (Favarel-Garrigues and Boget, 1968). Even in the apparent absence of infection, the temperature may be elevated as high as 41°C (Moene et al., 1969). Reawakening requires several hours in mild poisoning, 12–14 h in severe poisoning, and as much as 96 h in massive poisoning. The return of consciousness may be gradual but usually is sudden and may be accompanied by headache, stiffness, and weakness. Recovery without sequelae is the rule (Favarel-Garrigues and Boget, 1968; Tempé and Kurtz, 1972). Moene et al. (1969) reported a death caused by uncomplicated poisoning, but it is indicative of the toxicity

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of chloralose that the victim succeeded in his suicide with this compound only on his third attempt within a period of about 4 months. Death due to circulatory collapse occurred on the evening of the 5th hospital day. (c) Dosage Response Eleven patients, who were thought to have taken doses ranging from 640 to 2880 mg, all survived with appropriate treatments (Cornette and Franck, 1970). In another case involving chloralose intended as a rat poison, a dose thought to be 3000–4000 mg was survived (Boudouresque et al., 1966). Other reports of nonfatal dosages have fallen within the range of 2000–9000 mg, but one kind of rat poison came in 20-g packets and at least one person may have survived that dose. The dose in a fatal case was unknown (Gras et al., 1975; Moene et al., 1969). Tempé and Kurtz (1972) considered that toxic signs could result from 400 mg but that most cases resulted from ingestion of about 1000 mg (d) Laboratory Findings In two cases in which the dosage was known with considerable certainty, about 45% of the amount ingested was recovered from urine passed within the first 24 h, about 90% in the form of the glucose conjugate (Gras et al., 1975). The EEG picture in acute poisoning by chloralose is characteristic, involving slow waves and numerous spikes that usually are bilaterally symmetrical and synchronous. The pattern is changed dramatically by diazepam, which converts it to delta activity without the rapid rhythms commonly seen with that medication. The tracing always becomes normal, often within 24 h. The acute EEG record does not correspond to the myoclonic movements that, as observed visually or by EMG, are usually isolated, brief asymmetrical, and asynchronous (Boudouresque et al., 1966; Cornette and Franck, 1970; Tempé and Kurtz, 1972). (e) Treatment of Poisoning The myoclonic seizures respond to intravenous injection of 10 mg of diazepam, but this may have to be repeated once or twice. Recovery without sequelae is the rule (Boudouresque et al., 1966; Cornette and Franck, 1970; Tempé and Kurtz, 1972).

100.7.2  Norbormide 100.7.2.1  Identity, Properties, and Uses (a) Chemical Name 6-(-Hydroxy--2-pyridylbenzyl)-7-(-2-pyridylbenzylidene)-norbor-5-ene-2,3-dicarboximide is the chemical name. (b) Structure See Figure 100.6.

Hayes’ Handbook of Pesticide Toxicology

(c) Synonyms Norbormide (ANSI, BSI, ISO) is the common name for this compound. Trade names include Shoxin and Raticate. Code designations include McN-1,025 and S-6,999. The CAS registry number is 991-42-4. (d) Physical and Chemical Properties Norbormide has the empirical formula C33H25N3O3 and a molecular weight of 511.55. It is a white crystalline powder melting at 190–198°C. Its solubility at 30°C in ethanol is 14 mg/l; in chloroform, more than 150 mg/l; in diethyl ether, 1 mg/l; and in 0.1N HCl, 20 mg/l. Norbormide is stable at room temperature and to boiling. It is hydrolyzed by alkali and is noncorrosive. The structures of two diastereoisomers have been determined. They differ in the relative stereochemistry about the exocyclic double bond and the relative conformations of the aryl rings (Steel et al., 2004). (e) Analytical Method A thin-layer chromatographic/UV technique for the qualitative identification and quantitative determination of the major isomers of norbormide has been developed (Janicki et al., 1968). (f) History, Formulations, and Uses Norbormide was introduced in 1964 by the McNeil Laboratories, Inc. It is a selective rodenticide, lethal to rats but not to other rodent species. It usually is concentrated in prepared baits of cereal at 5–10 g/kg. Baits containing the compound should contain a warning dye. (g) Basic Findings Norbormide shows a remarkable selectivity both in toxicity and in pharmacological effect. Oral LD50 values for both wild and domestic Norway rats ranged from 5.3 to 15.0 mg/kg (Greaves, 1966; Niu, 1970; Roszkowski, 1965; Roszkowski et al., 1964). Corresponding values for roof rats and Hawaiian rats were 52 and about 10 mg/kg, respectively. Oral LD50 values were much higher in other rodents and lagomorphs: for example, hamster, 140 mg/kg; guinea pig, 620 mg/kg; mouse, 2250 mg/kg; and rabbit, about 1000 mg/kg. The oral toxicity was low in all other species tested; in the dog, cat, monkey, sheep, pig, and chicken, no effect was detectable at 1000 mg/kg (Niu, 1970; Roszkowski, 1965; Roszkowski et al., 1964). Rats given an overdose of norbormide died within 15 min to 4 h. At first, the animals assumed a hunched position. Later there was locomotor impairment due to weakness but not paralysis of the hind legs. Struggling, labored breathing and, in some instances, a mild convulsion preceded death (Roszkowski et al., 1964). Many analogues have been studied and none was found as toxic to rats as norbormide (Poos et al., 1996).

Chapter | 100  Rodenticides

Dogs survived daily doses of norbormide corresponding to a dietary level of 10,000 ppm for 15–60 days, but they lost appetite and looked ill. Dogs tolerated a dosage corresponding to a dietary level of 1000 ppm for 60 days without ill effect (Roszkowski et al., 1964). Even in laboratory rats, susceptibility to the compounds was greatly reduced when it was mixed in the diet rather than being given by stomach tube. This may be explained in part by tolerance. The oral LD50 determined after a 1-day rest period was less than doubled in rats pretreated for 1–7 days at the rate of 2 mg/kg/day. The degree of tolerance was small, but it was statistically significant. When the rest period was 5 days, no tolerance remained (Roszkowski, 1965). In any event, primary bait refusal can be a very serious problem in the use of norbormide (Maddock and Schoff, 1967). Mammals other than those in the genus Rattus, as well as birds, are resistant to the compound. For example, in the hamster, the most susceptible non-rat species, the LD50 is 140 mg/kg bw, while in cats, dogs, pigs, monkeys, mice, birds, etc., 1000 mg/kg bw produces neither death nor abnormal manifestations (Roszkowski et al., 1965). (h) Mode of Action and Cause of Death Some sex and species differences in susceptibility could be explained by differences in metabolism or absorption. For example, the oral LD50 values for male and female laboratory rats were 5.3 and 15.0 mg/kg, respectively, although the intravenous values (0.65 and 0.63 mg/kg, respectively) did not differ significantly. The very low susceptibility of mice to oral doses (LD50, 2250 mg/kg) was due in part to poor absorption, as indicated by the intraperitoneal LD50 of only 390 mg/kg. However, differences in absorption could not account for many observed species differences. For example, anesthetized dogs showed no detectable response to intravenous injection at the rate of 40 mg/kg. Species differences in response of the peripheral blood vessels to norbormide seemed to account for most of the observed species differences in its toxicity. The compound caused an extreme, irreversible vasoconstriction in laboratory rats, and this was considered the cause of death. The effect was demonstrated by direct observation of flow experiments, or both, in the ear, eye, skin, mesentery, and heart and undoubtedly occurred in other organs. It resulted from either systemic or appropriate local administration. Only vessels of relatively small caliber were visibly constricted. Spiral rat aortic segments and duodenal strips did not respond to norbormide. The mechanism of action was best demonstrated in the heart. Isolated myocardial strips showed no loss of contraction or responsiveness when norbormide was added to the bath. However, when norbormide was injected into the aorta of isolated rat hearts, the coronary flow rate decreased greatly and the heart slowed and developed arrhythmia. The effects were not inhibited or reversed by sodium nitrite or other vasodilators.

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Following vasoconstriction and presumably as a result of vascular injury, an erythematous but not typically inflammatory lesion began to develop in the rat skin 6 h after intradermal injection of 0.1 ml of 0.1% solution. The lesion became maximal in 24–48 h. Sometimes an area of central necrosis was observed. A concentration of 0.01% produced the effect only inconsistently. Except in rats, vasoconstriction was not seen even at high-dosage levels. Why rats respond differently remains obscure (Roszkowski, 1965). A number of the observations just discussed were confirmed by Niu (1970), who reached the same general conclusion regarding the importance of vasoconstriction in the rat. An oral or intraperitoneal dosage of 1020 mg/kg caused a doubling of blood glucose levels and a decrease of liver and muscle glycogen when coma began 0.5–2 h after treatment. The same dosage had no effect on the glucose levels of two strains of mice, nor did it produce illness. Insulin counteracted the hyperglycemic effect or norbormide in rats but did not protect against toxic manifestations and death, suggesting that the hyperglycemia is secondary (Patil and Radhakrishnamurty, 1973). The effects and mechanism of action were further investigated and the findings were reported in several publications. Experimental results obtained in an in vitro system (norbormide-sensitive rat caudal arteries and smooth muscle cells derived from rat mesenteric arteries and nonsensitive rat aorta) suggest that norbormide elicits its tissue and species selective vasoconstrictor effect by stimulating phospholipase C–protein kinase C pathway and calcium influx through verapamil-sensitive and -insensitive calcium channels (Bova et al., 2001a,b; Fusi et al. 2002). Cavalli et al., (2004) have shown that rat veins represented the main target for norbormide contractile action, that in both arteries and veins the norbormide-induced contractions were inversely related to the caliber of the vessels, that these contractions were mediated by the same mechanisms in both arteries and veins, and that in norbormidecontracted arteries the drug activates both contractile and relaxing mechanisms. Ricchelli et al. (2005) have contributed evidence of mechanistic species differences between the norbormide sensitive rats and the insensitive guinea pigs and mice. The differences are at the mitochondrial level where norbormide stimulates the opening of the mitochondrial permeability transition pores in various organs of the rat but not of guinea pigs and mice. In addition, norbormide induces rat-specific changes in the fluidity of lipid milieu of the internal mitochondrial membranes.

100.7.2.2  Toxicity to Humans In a study that has been cited for many years but apparently was first published by Hayes (1975), Dr. Kazwya Kamoya of the Showa Medical School in Tokyo administered

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norbormide to volunteers orally at doses ranging from 20 to 300 mg. No sign or symptom was produced. It was concluded that body temperature and blood pressure decreased slightly and temporarily following the larger dose. Actually, the largest fall in temperature observed at any dose was 0.7°C, and this occurred after doses of 20–80 mg, whereas the largest fall after a dose of 120 mg or more was only 0.3°C. Systolic (but not diastolic) blood pressure possibly fell after 120 mg of norbormide and certainly fell after larger doses. The largest decreases recorded were from 132/76 to 100/74 after 200 mg and from 120/80 to 96/80 after 300 mg. The lowest values were measured 1 h after ingestion, and the values were essentially normal at 2 h in each instance. It seems unlikely that poisoning by norbormide will occur. If it does, treatment must be symptomatic.

100.7.3  Bromethalin 100.7.3.1  Identity, Properties, and Uses (a) Chemical Name The IUPAC chemical name is ,,-trifluoro-N-methyl4,6-dinitro-N-(2,4,6-tribromophenyl)-o-toluidine and the CAS chemical name is N-methyl-2,4-dinitro-N-(2,4,6-tribromophenyl)-6-(trifluoromethyl)benzenamine. (b) Structure See Figure 100.7. (c) Synonyms Code designations include EL 614, Lilly 126714, OMS 3020. The CAS registry number is 63333-35-7. (d) Physical and Chemical Properties Bromethalin has the empirical formula C14H7Br3F3N3O4 and a molecular weight of 577.9. It forms pale yellowish crystals with a specific gravity of 1.36 and a melting point of 151°C. Its water solubility is low: 0.002 mg/l at 20°C and its vapor pressure at 25°C is 0.013 mPa. The octanol/water partition coefficient (Log P) at ph 7 and 20°C is 7.68. (e) History, Formulations, and Uses Bromethalin is a nonanticoagulant rodenticide of the diphenylamine chemical family. Its properties were discovered in 1979. It is a single-feeding rodenticide with efficacy against coumarin-resistant rodents. Baits contain either 0.01 or 0.005% active material (Jackson et al., 1982). Br CF3 N

O2N

NO2

Br

Br

Figure 100.7  Chemical structure of bromethalin.

100.7.3.1  Toxicity to Laboratory Animals The LD50 values of bromethalin for male and female Wistar rats (Rattus norvegicus) are 2.01 and 2.46 mg/kg bw, respectively, whereas the LD50 is 6.60 mg/kg bw for male and female wild rats (Rattus rattus). In female mice (Mus musculus), the oral LD50 is 8.13 mg/kg bw and 5.25 mg/kg bw in males (Jackson et al., 1982). In dogs, the oral LD50 is 4.7 mg/kg bw, it is 1.8 mg/kg bw in cats, 5.0 mg/kg bw in monkeys, and 13 mg/kg bw in rabbits. Several subchronic studies have been conducted in rats. The most characteristic feature elicited by the studies is a vacuolation of the white matter of the brain and the spinal cord. This spongy degeneration has been shown to be partially reversible after an off-treatment recovery period of 2 and 4 weeks. After 8 weeks postexposure, the white matter histological lesions were no longer apparent. The clinical signs observed during those studies included leg weakness and depressed motor function. A NOEL of 25 g/kg bw was demonstrated in both rats and dogs (Spaulding and Spannring, 1988). In a group of four cats treated with a single oral dose of 1.5 mg/kg bw, a neurotoxic syndrome developed including ataxia, focal motor seizures, vocalization, decerebrate posture, recumbency, depression, and semi-coma. Survival time ranged from 48 h to 97 h. Typical spongy degeneration of the white matter was seen in the cerebrum, cerebellum, brain stem, and spinal cord as well as in the optic nerve of all treated cats. Ultrastructural findings included separation of myelin lamellae with formation of coalescent intramyelinic vacuoles and pronounced cytosolic edema of astrocytes and oligodendroglial cells (Dorman et al., 1992). A dose of 6.25 mg/kg bw of bromethalin in dogs induced a toxic syndrome characterized by hyperexcitability, tremors, seizures, depression, and death within 15–63 h after dosing. Histologic lesions included diffuse white matter spongiosis in all five treated dogs, mild microgliosis, and optic nerve vacuolization (Dorman et al., 1990). Bromethalin toxicosis produced acute and chronic electroencephalographic (EEG) changes including spike and spike-and-wave patterns, high-voltage/slow wave activity, and marked voltage depression (U.S. EPA Reregistration Eligibility Decision Rodenticide Cluster, 1998).

100.7.3.2  Toxicity to Nontarget Species Bromethalin is toxic to most nontarget species tested. In birds (adult quails), the single oral LD50 is 4.6 mg/kg bw, whereas it is 210 ppm when administered in the feed. It is also toxic to aquatic species: the LC50 in bluegill: 120 ppb, in trout: 80 ppb but 30 ppb and 27 ppb in daphnia (Jackson et al., 1982). Bromethalin poisoning is not uncommon in domestic animals. In 1987, bromethalin-based rodenticides were the sixth most commonly ingested rodenticide by animals as reported to the Illinois Animal Poison Information Center (IAPIC), accounting for 148 calls compared to brodifacoum 4061 calls, warfarin 488 calls, cholecalciferol 362 calls, and bromadiolone 225 calls (Dorman et al., 1990).

Chapter | 100  Rodenticides

100.7.3.3  Metabolism, Tissue Distribution, and Mechanism of Action Several studies have been conducted to investigate the mechanism of action, metabolism, excretion, and mechanism of action of bromethalin. Following oral administration, bromethalin is metabolized to desmethyl-bromethalin by the hepatic mixed function oxidase. This metabolite uncouples oxidative phosphorylation in mitochondria. The sodium/potassium gradient is disturbed and fluids build up in cells. In the central nervous system, increased pressure induces demyelination, which in turn disturbs the nerve impulse transmission. In lethally poisoned animals, these disturbances lead to inadequate transmission to the respiratory centers, resulting in respiratory arrest and death. The entire process is dose-dependent and death generally occurs from 12 to 72 h following ingestion (Cherry et al., 1982; van Lier and Cherry, 1988). Studies with 14C-bromethalin have provided information leading to the determination of a half-life of approximately 135 h. After oral administration, bromethalin is primarily distributed to the fat, liver, gastrointestinal tract, adrenals, ovaries, and thyroid. Excretion is essentially through the feces (van Lier and Cherry, 1988; van Lier and Ottosen, 1981).

100.7.3.4  Toxicity to Humans Although accidental poisoning in nontarget/domestic animals, dogs, and cats but essentially dogs are not uncommon, reports of human poisonings remain scarce. One fatal intentional ingestion of bromethalin-based rodenticide has been published (Pasquale-Styles et al., 2006). A 21-year-old mentally disturbed male was admitted to the hospital the day after ingesting a bromethalincontaining rodenticide (equivalent to approximately 17 mg of active ingredient or 0.33 mg/kg bw). The main clinical and pathological features were altered mental status, increased cerebrospinal fluid pressure, and cerebral edema. He died 7 days after exposure. Diffuse histologic vacuolation of the white matter of the central nervous system was observed at autopsy. The demethylated form of bromethalin was detected in the liver and brain by gas chromatography-mass spectrometry (GC-MS). (a) Treatment of poisoning On the basis of the metabolism of bromethalin, Dorman et al. (1990) have tried to develop an antidote. Bromethalin is N-demethylated by mixed function oxidase system to its desmethyl-bromethalin, which is a 10- to 1000-fold more potent uncoupler of oxidative phosphorylation than bromethalin (van Lier and Cherry, 1998). The apparent lack of toxicity of bromethalin observed in guinea pigs may be partially due to the relative deficiency of this species in N-demethylase activity. Therefore a compound that would inhibit the conversion of bromethalin to its desmethyl metabolite would be a good antidote candidate.

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Aminophylline was selected by the authors based on their selection criteria. When administered to rats given a uniformly lethal dose of bromethalin, aminophylline was not associated with a statistically significant increase in survival time. Bromethalin administration produced the known toxic syndrome in all treated rats. At this point, aminophylline does not appear to be a valid antidote to bromethalin poisoning. Treatment is supportive. (b) Analytical Methods Bromethalin and its main metabolite desmethyl-bromethalin can be detected in baits, biological fluids, and tissues by HPLC with UV detection or by HPLC coupled with atmospheric pressure chemical ionization (APCI)mass spectrometry (MS) (Mesmer and Flurer, 2001).

100.7.4  Banned Compounds Tetramethylenedisulfotetramine has never been registered in the United States where it is banned for import and is not registered anywhere else in the world. It was banned in China in 1984 but is apparently still manufactured and sold and obviously illegally exported. It is an odorless, tasteless, and water-soluble white crystalline powder. Its chemical formula is C4H8N4O4S2 and its molecular weight is 240. It acts as a gamma-amino butyric acid (GABA) antagonist that binds noncompetitively and irreversibly to the GABA receptors on the neuronal cell membrane and blocks chloride channels. It is a potent convulsant. Death usually occurs within hours of ingestion. The LD50 of tetramethylenedisulfotetramine in mammals is 0.1–0.3 mg/kg, and 7–10 mg (total dose) is considered to be a lethal dose for human beings. The first case of known human poisoning that occurred in the United States in 2002 was reported by Barrueto et al. (2003). Cases have also been reported from China (Chau et al., 2005; Lu et al., 2008; Poon et al., 2005). This compound’s properties and its toxic clinical and pathological manifestations have recently been extensively reviewed (Whitlow et al., 2005).

100.7.5  On-going Research for New Chemical Families with Rodenticidal Properties New naturally occurring substances are presently being investigated in order to tentatively identify new families that could be developed as rodenticides. Very few data have been published. A number of examples are presented here. (a) Natural Coumarins Extracted from Citrus bergamia The bergamote (Citrus aurantium ssp. bergamia) is a small and roughly pear-shaped citrus fruit originating and grown

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mainly in Calabria, Italy. It is also cultivated in Argentina, Brazil, the United States (Georgia), and India. An essence extracted from the aromatic skin of this sour fruit is used in the flavoring and perfume industries. Ethanol and petroleum ether extracts of the fruit pulp have been administered to Rattus norvegicus and to Rattus rattus (El-Naggar and Mikhail, 1999). Two natural coumarins were isolated from the petroleum ether extract: bergapen and coumourrayin. Norway rats (R. norvegicus) were more susceptible to both compounds than roof rats (R. rattus). (b) Extract from Seeds and Bark of Nerium indicum Crude extract of Nerium indicum is highly toxic to Bandicota bengalensis. A dose of 12.5 ml/kg bw killed 100% of the group of animals (Saravan et al., 2004). N. indicum contains a large number of toxic glycosides of which oleandrin is the most important one (0.018–0.425%). Numerous cases of human poisonings are known, essentially accidental ingestion in children (Gaillard et al., 2001). The rodenticidal performance of extract is still being investigated in several countries. N. indicum leaf and bark extracts are used as molluskicides in India. (c) Mustard Seed Powder (Brassica hirta) A rodenticide containing a mixture of powder of mustard seeds (Brassica hirta) and sodium -olefinsulfonate is registered in the United States and Canada for the control (and not for the eradication) of ground squirrels (Spermophilus elegans). The product is prepared as a water suspension that is injected with a special olefinsulfonate foam-making nozzle into the burrows. The mode of action is by asphyxia of animals breathing the foam. Mustard seed powder induces respiratory irritation that speeds up the lethal effect. Both the American and the Canadian authorities have determined that the compound did not pose any particular risk to human health or to the environment when used according the label’s instructions.

Conclusion In consideration of the vast impact of rodents’ activities on public health throughout the world, but essentially in the developing countries, the usefulness of rodenticides to control their populations is unquestionable. Rodents of all sorts compete with humans for food and are responsible for significant economic losses in countries where food is already scarce and hard to produce. In some countries, they can cause starvation. Rodents are also hosts for a relatively large number of severe human diseases such as endemic rickettsiosis, leishmaniasis, spirochetosis, tularemia, leptospirosis, tick-borne encephalitis, and listeri­ osis. In addition, they do a variety of other types of damage to buildings such as gnawing wires and pipes.

Effective permanent population control is technically difficult. Rodenticides are essential among other means to fight against rodents. Research and development has been active worldwide in searching for substances that are effective against rodents while having an acceptable margin of safety for humans. Products from many chemical families have been developed. Their mechanisms of action differ. Some of them are very toxic to humans and must be used by trained professionals but, most often, this is far from being the case in developing countries where accidental poisonings are frequent and antidotes for these compounds are not available. Others have a mode of action that requires repeated dose absorptions to be effective in rodents. This the case of anticoagulants. Several substances have been developed and marketed in this category and are now widely used. They are active (and toxic) at very low doses that must be repeated over a few days. In humans, apart from misuse or careless professional exposure, the risk of repeated absorption is limited. In addition, an antidote exists and allows for effective treatment (vitamin K). Attempts have been made to identify and develop socalled natural substances extracted from plants but so far without real success.

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Gunther, R., Felice, L. J., Nelson, R. K., and Franson, A. M. (1988). Toxicity of vitamin D3 to dogs. J. Am. Vet. Med. Assoc. 193, 211–214. Hagan, E. C., and Radomski, J. L. (1953). The toxicity of 3-(acetonylbenzyl)-4-hydroxycoumarin (warfarin) to laboratory animals. J. Am. Pharm. Assoc. 42, 379–382. Hall, J. G., Pauli, R. M., and Wilson, K. M. (1980). Maternal and fetal sequelae of anticoagulation during pregnancy. Am. J. Med. 68, 122–140. Hanriot, M., and Richet, C. (1897). Chloralose. Arch. Int. Pharmacodyn. 3, 191–211 [in French]. Hantson, P., Bernard, A., and Hermans, C. (2008). Kinetics and determinants of the changes of CC16, a lung secretory protein in a rat model of toxic lung injury. Clin. Toxicol. 46(3), 230–238. Harris, L. J. (1955). “Vitamins in Theory and Practice,” 4th ed. Cambridge Univ. Press, London/New York. Harrisson, J. W. E., Ambrus, J. L., and Ambrus, C. M. (1952a). Fluoroacetate (1090) poisoning. Ind. Med. Surg. 21, 440. Harrisson, J. W. E., Ambrus, J. L., Ambrus, C. M., Rees, E. W., Peters, R. H., Reese, L. C., and Baker, T. (1952b). Acute poisoning with sodium fluoroacetate (compound 1080). J. Am. Med. Assoc. 149, 1520–1522. Hartley, D., and Kidd, H. (1983). “The Agrochemicals Handbook.” The Royal Society of Chemistry, Nottingham, UK. Hatakeyama, H., and Shigei, T. (1971). Fine-structured changes of alveolar walls in fibrin-induced, so-called neurogenic pulmonary edema of the rat: Comparative representation with adrenaline and ANTUinduced edema. Jpn. J. Pharmacol. 21, 673–675. Havill, A. M., Gee, M. H., Washburne, J. D., Premkumar, A., Ottavianon, R., Flynn, J. J., and Spath, J. A. (1982). Alpha-naphthylthiourea produces dose dependent lung vascular injury in sheep. Am. J. Physiol. 243, 505–511. Hayes, W. J. Jr. (1975). “Toxicology of Pesticides.” Williams & Wilkins, Baltimore. Hayes, W. J. Jr., and Gaines, T. B. (1950). Control of Norway rats with residual rodenticide warfarin. Public Health Rep. 65, 1537–1555. Hayes, W. J. Jr., and Vaughn, W. K. (1977). Mortality from pesticides in the United States in 1973 and 1974. Toxicol. Appl. Pharmacol. 42, 235–252. Herken, H. (1971). Antimetabolic action of 6-aminonicotinamide on the pentose phosphate pathway in the brain. In “Mechanisms of Toxicity,” (W. N. Aldridge, ed.). Macmillan, London. Hermans, C., Lesur, O., Weynand, B., Pieters, Th., Lambert, M., and Bernard, A. (1998). Clara cell protein (CC16) in pleural fluids. A marker of leakage through the visceral pleural. Am. J. Resp. Care Med. 157(3), 962–969. Hirsh, J., Dalen, J. E., Anderson, D. R. et al. (2001). Oral anticoagulants: mechanism of action, clinical effectiveness and optimal therapeutic range. Chest 119(suppl), 8S–21S. Holbrook, AM., Pereira, J. A., Labiris, R., McDonald, H., Douketis, J. D., Crowther, M., and Wells, P. S. (2005). Systematic overview of warfarin and its drug and food interactions. Arch. Intern. Med. 165(10), 1095–1106. Hollinger, B. R., and Pastoor, T. P. (1993). Case management and plasma half-life in a case of brodifacoum poisoning. Arch. Intern. Med. 153, 1925–1928. Holmes, R. W., and Love, J. (1952). Suicide attempt with warfarin, a bishydroxycoumarin-like rodenticide. J. Am. Med. Assoc. 148, 935–937. Hosek, R. W., and Love, J. (1952). Mathematical evaluation of the course of body temperature in mice after the administration of sodium fluoro­ acetate. Physiol. Bohemoslov. 19, 129–133.

Chapter | 100  Rodenticides

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Soubiron, L., Hantson, P., Michaux, I., Lambert, M., Mahieu, P., and Pringot, J. (2000). Spontaneous haemoperitoneum from surreptitious ingestion of a rodenticide. Eur. J. Emerg. Med. 7(4), 305–307. Spaulding, S. R., and Spannring, H. (1988). Status of bromethalin outside the United States. Proc. 13th Vertebr. Pest Conf., Univ. Nebraska, Lincoln (A.C. Crabb, and R.E. Marsh, eds.) University of Calif., Davis, 13:64–69. Spielmann, H., Meyer-Wendecker, R., and Spielmann, F. (1973). Influence of 2-deoxy-D-glucose and sodium fluoroacetate on respiratory metabolism of rat embryos during organogenesis. Teratology 7, 127–134. Spiller, H. A., Gallenstein, G. L., and Murphy, M. J. (2003). Dermal absorption of a liquid diphacinone rodenticide causing coagulopathy. Vet. Hum. Toxicol. 45(6), 313–314. Spurr, E. B., Maitland, M. J., Taylor, G. E., Wright, G. R. G., Radford, C. D., and Brown, L. E. (2005). Residues of brodifacoum and other anticoagulant pesticides in target and non-target species, Nelson Lakes National Park, New Zealand. N. Z. J. Zool. 32, 237–249. Steel, P. J., Brimble, M. A., Hopkins, B., and Rennison, D. (2004). Two stereo­isomers of the rat toxicant norbormide. Acta Cryst. 60, 374–376. Steinberger, E., and Sud, B. N. (1970). Specific effects of fluoroacetamide on spermiogenesis. Biol. Reprod. 2, 369–375. Stevenson, R. E., Burton, M., Ferlanton, G. J., and Taylor, H. A. (1980). Hazards of oral anticoagulants during pregnancy. J. Am. Med. Assoc. 243, 1549–1551. Stone, W. B., Okoniewski, J. C., and Stedelin, J. R. (1999). Poisoning of wildlife with anticoagulant rodenticides in New York. J.Wildlife Dis. 35(2), 187–193. Stowe, C. M., Metz, A. L., Arendt, T. D., and Schultman, J. (1983). Apparent brodifacoum poisoning in a dog. J. Am. Med. Assoc. 182, 817–818. Suttie, J. W., ed. (1979). “Vitamin K Metabolism and Vitamin KDependent Proteins.” University Park Press, Baltimore. Talcott, P. A., Mather, G. G., and Kowitz, E. H. (1991). Accidental ingestion of a cholecalciferol-containing rodent bait in a dog. Vet. Hum. Toxicol. 33, 252–256. Tarrant, K. A., and Westlake, G. E. (1984). Histological technique for the identification of poisoning in wildlife by the rodenticide calciferol. Bull. Environ. Contam. Toxicol. 32, 175–178. Tejani, N. (1973). Anticoagulant therapy with cardiac valve prothesis during pregnancy. Obstet. Gynecol. 42, 785–793. Tempé, J. D., and Kurtz, D. (1972). Acute chloralose poisoning. Concours Med. 94, 801–813 [in French]. Terneu, S., Verhelst, D., Thys, F., Ketelslegers, E., Hantson, P., and Wittbole, X. (2003). An unusual cause of abdominal pain. Acta Clin. Belg. 58(4), 241–244. The Merck Veterinary Manual, 9th ed. (2008). Kahn, C.J. ed. Merck & Co. Whitehouse Station, N.J. Thierry, M. J., Hermodson, M. A., and Suttie, J. W. (1970). Vitamin K and warfarin distribution and metabolism in the warfarin-resistant rat. Am. J. Physiol. 219, 854–859. Thompson, R. D., Mitchell, G. C., and Burns, R. J. (1972). Vampire bat control by systemic treatment of livestock with an anticoagulant. Science 177, 806–807. Thompson, A. R. (1996). Molecular hemostatic variant enhances warfarin toxicity. J. Clin. Invest. 98(7), 1508. Toda, T., Ito, M., Toda, Y., Smith, T., and Kummerov, F. (1985). Angiotoxicity in swine of a moderate excess of dietary vitamin D3. Food Chem. Toxicol. 23, 585–592. Tokavera, T. G., Turov, I. S., and Alekseyev, A. N. (1971). The action of fluoroacetamide on albino mouse fecundity (preliminary report). Zh. Mikrobiol. Epidemiol. Immunobiol. 48, 24–26 [in Russian].

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Tomiya, K., Shimoda, R., Kakihara, Y., and Kanda, M. (1976). Studies on extraction of toxic materials from organs of victims of fatal intoxication by organofluorine pesticides. Jpn. J. Leg. Med. 29, 237–238 [in Japanese]. Tomlin, C. (1994). “The Pesticide Manual,” 10th ed. British Crop Protection Council. Farnam, Surrey, United Kingdom. Tourtellotte, W. W., and Coon, J. M. (1951). Treatment of fluoroacetate poisoning in mice and dogs. J. Pharmacol. Exp. Ther. 101, 82–91. Travis, S. F., Warfield, W., Greenbaum, B. J., Moloskier, M., and Siegel, J. E. (1993). Spontaneous hemorrhage associated with accidental brodifacoum poisoning in a child. J. Pediatr. 122(6), 982–984. Ullrich, V., and Staudinger, H. (1968). Metabolism in vitro of warfarin by enzymic and nonenzymic systems. Biochem. Pharmacol. 17, 1663–1669. United States Environmental Protection Agency. (1998). Reregistration Eligibility Decision. Rodenticide Cluster. 296 p. EPA738-R-98-007. Washington, D.C. July 1998. Valchev, I., Binev, R., Yordanova, V., and Nikolov, Y. (2008). Anticoagulant rodenticide intoxication in animals – A review. Turk. J. Vet. Anim. Sci. 32(4), 237–243. Vandenbroucke, V., Bousquet-Melou, A., DeBaker, P., and Croubels, S. (2008). Pharmacokinetics of eight anticoagulant rodenticides in mice after single oral administration. J. Vet. Pharmacol. Therap. 31, 437–445. Van Driel, D., Wesseling, J., Sauer, P. J. J., Touwen, B. C. J., van der Veer, E., Hugo, S.A., and Heymans, H. S. A. (2002). Teratogen update: Fetal effects after in utero exposure to coumarins. Overview of cases, follow-up findings and pathogenesis. Teratology 66, 127–140. van Lier, R. B. L., and Ottosen, L. D. (1981). Studies on the mechanism of toxicity of bromethalin, a new rodenticide. Toxicologist 1(1), 114 [Abstract]. van Lier, R. B. J., and Cherry, L. D. (1988). The toxicity and mechanism of action of bromethalin: a new single-feeding rodenticide. Fund. Appl. Toxicol. 11(4), 664–772. Varon, M. L., and Cole, L. J. (1966). Hemopoietic colony-forming units in regenerating mouse liver suppression by anticoagulants. Science 153, 643–644. Vasquez, A. G., and Rodriguez, M. A. (2000). Sindromo hemorragiparo por esposicion a raticida. Communicacion de un caso. Rev. Med. Chile 128(6), 647–649. Vaughan, E. D. Jr., Moore, R. A., Warren, H., Moler, D. N., and Gillenwater, J. Y. (1969). Skin necrosis of genitalia and warfarin therapy. J. Am. Med. Assoc. 210, 2282–2283. Vik, T., Try, K., Thelle, D. S., and Forde, O. H. (1979). Tromso heart study: Vitamin D metabolism and myocardial infarction. Br. Med. J. 2, 176. Vindenes, V., Karinen, R., Hasvold, I., Bernard, J. P., Morland, J. G., and Christophersen, A. S. (2008). Bromadiolone poisoning: LC-MS method and pharmacokinetic data. J. Forensic Sci. 53(4), 993–996. Vogel, J. J., de Moerloose, P., Bouvier, C. A., Gaspoz, J., and Riant, P. (1988). Anticoagulation prolongée lors d’une intoxication à la chlorophacinone. Schweiz Med. Wochenschr. 118, 1915–1917. Watt, B. E., Proudfoot, A. T., Bradberry, S. M., and Vale, J. A. (2005). Anticoagulant rodenticides. Toxicol. Rev. 24(4), 259–269. Ward, J. C. (1945). Rodenticides – Present and future. Soap Sanit. Chem. 21(9), 117 119, 127. Ward, J. C., and Spencer, D. A. (1947). Notes on the pharmacology of sodium fluoroacetate-compound 1080. J. Am. Pharm. Assoc. Sci. Ed. 36, 59–62. Wells, P. S., Holbrook, A. M., Crowther, N. R., and Hirsh, J. (1994). Interactions of warfarin with drugs and food. Ann. Intern. Med. 121(9), 676–683. Wenz, C. (1984). New chemicals under fire. Nature (London) 1, 741.

Chapter | 100  Rodenticides

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Chapter 101

Toxicology and Safety Evaluation of the New Insect Repellent Picaridin (Saltidin) G.K. (Ghona) Sangha Lanxess Corporation, Pittsburgh, Pennsylvania

101.1  Introduction Picaridin is a new generation of customized active ingredient specifically designed to repel a variety of arthropods and is marketed under the name Saltidin (registered trade name of Saltigo GmbH, LANXESS Group). It was developed by Bayer as an alternative to DEET and is now owned by LANXESS Corporation (previously a Division of Bayer Corporation). It was tailor-made, based on the hypothesis that the repellent effect is triggered by the action of a given substance on specific olfactory receptors of the arthropod. Molecular modeling techniques were utilized during the development process, which allows for the three-dimensional construction and mapping of molecules. Existing repellent products were altered at specific sites where an interaction with an arthropods’ receptor was anticipated. More than 800 substances were synthesized and screened for efficacy, cosmetic properties and safety. Picaridin (laboratory name KBR 3023) represented the best compromise of all the required properties for an ideal repellent. Picaridin showed the best performance regarding efficacy against a variety of arthropods (Boeckh et al., 1996) and had the most desired attributes regarding safety as well as compatibility with skin and plastic materials. The report of the fourth WHOPES (WHO Pesticide Evaluation Scheme) Working Group meeting in December, 2000 (WHO, 2000) concluded that Picaridin has a good safety profile and cosmetic properties, ���������������������������������������������������������� and recommended it as the repellent of choice for malaria prevention. Alone �������������������������������������������������� or in typical formulations, it does not significantly attack common household materials including plastics, coatings, foils, and varnishes.

developed under the laboratory code name KBR 3023 with the registered trade name Bayrepel and was subsequently sold under the brand name Autan. The chemical name (CAS) for Picaridin is 1-piperidine­ carboxylic acid, 2-(2-hydroxy-ethyl), 1-methylpropylester. However, the INCI (International Nomenclature of Cosmetic Ingredients) name was given as hydroxy ethyl isobutyl piperidine carbamate. The common name, Picaridin, was rejected by ISO (International Organization for Standards) as it was not considered a pesticide. The common name Picaridin was also rejected by WHO/INN (World Health Organization/International Non-proprietary Name) but the common name, Icaridin, was accepted by WHO/INN. The structural formula is presented in Figure 101.1. Picaridin has high stability toward light, oxidation, water, and sweat. It has good compatibility with skin and mucous membranes, is compatible with plastic materials and has low skin penetration. Key physical-chemical of Picaridin are shown in Table 101.1.

101.2.2  Mode of Action In general, repellent activity is thought to be the result of any number of physiological and biochemical events. Mosquito repellency is thought to be due to blocking of the lactic acid receptors resulting in “losing” the host (Peterson and Coats,

HO

101.2  General overview 101.2.1  Chemistry During product development, Picaridin (U.S. registered name) was identified by several names. The current common name is Picaridin, and the current trade name Saltidin. It was Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

O

N

H3C CH3

O

Figure 101.1  Structural formula.

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Hayes’ Handbook of Pesticide Toxicology

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Table 101.1  Key Physical-chemical Properties of KBR 3023 Parameter (units)

Value

CAS No.

119515-38-7

Chemical name

1-piperidine carboxylic acid, 2-(hydroxy-ethyl), 1-methylpropylester

Molecular formula

C12H23NO3

Physical state

Liquid

Molecular weight (g/mol)

229.3

Color

Colorless 3

Density (g/cm at 20°C)

1.04

Solubility in water (at 20°C)

8.6 g/l in unbuffered 8.2 g/l in buffered

Dissociation constant

No acidic or basic properties in aqueous solutions

Solubility in organic solvents (at 20°C)

250 g/l

Flashpoint (°C)

142

Freezing point (°C)

170

Boiling point (°C)

272

Vapor pressure (kPa at 20°C)

3.40E-04

2001). Electrophysiological studies on the insects’ olfactory receptor organs reveal that certain cell types, which are not involved in perception of the attractive odorants, respond to Picaridin. As soon as Picaridin is presented together with an attractant, a new input is activated in the nervous system, which adds to the input from other receptors activated by the attractant. This new overall pattern clearly differs from that elicited by the attractant, so that the insect is no longer able to detect the attractant. The specificity and mode of action of Picaridin was investigated by experiments exploring second-messenger responses of male cockroaches. Picaridin induced a rapid increase in the concentration of inositol triphosphate in a dose-dependent and tissue-specific manner; other second-messenger systems were not affected. These observations suggest that Picaridin may act via subsets of G-protein-coupled receptors in sensory neurons. The repellent effect of the active substance starts immediately after application on the skin and develops full performance within a few minutes (Boeckh et al., 1996).

101.2.3  Effectiveness Against   Disease Vectors Biting and blood feeding by the arthropod pests can cause annoyance, blood loss, allergic reaction, and may be the

means by which people get infected with pathogenic organisms. Arthropod-borne pathogens are known to occur in North America and throughout the world. The most important are malaria, West Nile Virus, several types of viral encephalitis (Western Equine, Eastern Equine, and St. Louis) transmitted by mosquito species, Lyme Disease transmitted by tick species and leishmaniasis by the sand fly. Picaridin containing insect repellent formulations (cream, pump spray, aerosol, or wipes) containing 5–20% Picaridin have broad spectrum efficacy and are highly effective against a variety of blood sucking arthropod pests, especially the primary pest, mosquitoes (Aedes aegypti, Aedes albopictus, Aedes increpitus, Aedes melanimon, Aedes nigromaculis, Aedes sierrensis, Aedes sticticus, Aedes vexans, Culex quinquefasciatus, Culex tarsalis, Ochlerotatus taeniorhynchus, Anopheles dirus, Anopheles freeborni, Anopheles franciscanus). The other arthropod pests that are effectively repelled by Picaridin include ticks (Ixodes ricinus, Ixodes scapularis), flies (Tabanus bovinus, Haematopota pluvialis, Stomoxys calcitrans, Chrysops relictus), gnats, chiggers, biting midges (Culicoides impunctatus), fleas, and sand flies (Phlebotomus). Based on broad-spectrum effectiveness to biting arthropod species, Picaridin products are intended for use by nonprofessional users of the general public including adults and children.

101.2.3.1  Effectiveness Numerous field and laboratory tests of various formulations of Picaridin using human volunteers and animals have been conducted in various parts of the world by different researchers, including studies sponsored by the producers of the product, World Health Organization, and the militaries of many countries including the U.S. and Australian armies. These tests have shown Picaridin to be a safe and effective repellent of numerous pests and effectiveness was dependent on species, population density, geographical region, and concentration of Picaridin in the formulation (Table 101.2). The 20% formulations provide 8-h protection against various arthropod species, including mosquitoes, ticks, biting flies, sand flies, and biting midge. In recent field studies conducted in the United States, formulations containing 20% Picaridin provided 12–14 h of protection against mosquitoes including those carrying West Nile Virus (Carroll, 2008a). A similar field test conducted with 15% formulation in combination with sunscreen showed an enhancement in the effectiveness against mosquitoes (Carroll, 2008b). In a laboratory test with ticks using the sunscreen combination effectiveness against ticks was slightly decreased (Carroll, 2008c). The report of the fourth WHOPES Working Group meeting in December, 2000 concluded that: “KBR 3023 was tested under temperate and tropical conditions against important disease vectors Aedes albopictus, Anopheles

Chapter | 101  Toxicology and Safety Evaluation of Picaridin

2221

Table 101.2  Effectiveness of Picaridin Against Various Disease Vectors in the Field Environment Species

Formulation (% Picaridin)

Protection time (hours)

Reference

Mosquitoes

20%

8–14

15%a 15% 10% 5%

11.7–13.3 10.3 4–8 4–7

Carroll (2008a,b); Yap et al. (1998, 2000); Frances et al. (2002, 2004); Barnard et al. (2002); Luepkes (2005)

20%

4– 8.2–8.7 9.7–11.8

Sixl (1993); Carroll (2008c); Todd (1997)

15%a 15% Flies

12% 7.5%

7 4–6

Nentwig (1997); Muelhofer (1993)

Sand fly

20% 10%b

7.7–10.3 8.8–9.1

Perrotey (2002)

Biting midge

20%

7–8

Mordue (1999); Carpenter (2005)

Fleas

20%

9

Nentwig (1998)

Ticks

a With b

sunscreen. With polyester acrylate.

gambiae and Culex quinquefasciatus and several pest mosquitoes, demonstrating excellent repellent properties comparable to, and often superior to those of the standard DEET.” KBR 3023 conferred more than 95% protection up to 6–7 h after application. At comparable doses, KBR 3023 showed significantly longer protection times than DEET against Anopheles gambiae complex malaria vectors. “KBR 3023 can be recommended as the repellent of choice for malaria prevention.”

101.3  Metabolism and toxicokinetics The absorption, distribution, metabolism and excretion of [14C]-Picaridin were studied in rats and human volunteers following dermal application. The dermal route was chosen instead of the oral route because Picaridin is used in dermally applied insect repellents and dermal absorption is the foreseeable route of systemic exposure. In the rat, the test compound was administered either intravenously as a single dose of 20 mg/kg body weight (bw) or dermally as single doses of 20 or 200 mg/kg bw. In addition, [14C]-Picaridin was applied dermally to rats at a dose of 20 mg/kg bw following 14 daily dermal exposures to unlabeled Picaridin. Rats in the i.v. study were sacrificed 2 days after injection, whereas rats in the dermal study were sacrificed immediately after the 7-day-exposure period.

These data indicate that renal excretion is the principal pathway for elimination of KBR 3023 from rats following either intravenous or dermal dosing. Nineteen metabolites were identified in urine and feces of both i.v. and dermally dosed rats (Figure 101.2). Analysis of excreta from i.v. dosed rats identified 61–78% of the dose in urine and 6% of the dose in feces. For the dermally dosed groups, analysis of urine and feces identified 24–53% of the dose, or 75–85% of the radioactivity recovered in excreta. With the exception of two metabolites M18 and M19 (each at 0.3% of dose) that were found only in feces, the metabolite profile and relative distribution of metabolites was the same in urine and feces. There was also no qualitative difference in the metabolite profile between dose groups and sexes. The metabolism of KBR 3023 in rats primarily involves oxidation of the 2-hydroxyethyl group to an acid to form metabolite M16, coupled with hydroxylation of the 1-methylpropyl group to form metabolites M8, M9, and M10. The other minor phase I metabolites result from hydroxylation of the piperidine ring (Ml–M4 and M7). Minor phase II metabol­ ites result from conjugation of glucuronic acid with parent (M14 and M15) or phase I metabolites (M5, M6, and M11–M13). The metabolic profile in the human was investigated from the dermal absorption study in male volunteers dosed dermally on a forearm with either neat [14C] KBR 3023 (98% purity) or [14C] KBR 3023 in ethanol (15% w/w) at

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Figure 101.2  Proposed metabolic pathway for KBR 3023 in rats.

a nominal dose level of 15 mg/person (0.625 mg/cm2) (Selim, 1994). Nearly all the absorbed radioactivity was excreted in the urine (1.66–3.76%) for both dose groups, with 94% of the radioactivity in urine being excreted within 24 h of initial dosing. Concentrations of radioactivity in plasma drawn from the contralateral arm were similar between the two dose groups over time. Radioactivity in plasma increased to a maximum at 8 h when the treatment site was washed and then declined rapidly, returning to background levels by 8 h after removal of the test substance. Biotransformation of [14C] KBR 3023 in humans primarily involved conjugation of KBR 3023 with glucuronic acid

through the 2-hydroxyethyl moiety to form metabolites M14 and M15 (Figure 101.2), which together accounted for 43.1% of the radioactivity in urine. Other major metabolites in urine included M5 (17.4%), another glucuronic acid conjugate that is hydroxylated in the 1-methyl-propyl moiety; M16 (8.5%), in which the 2-hydroxyethyl group is oxidized to an acid; M8 (6.2%), in which the 2-hydroxyethyl group is oxidized and the 1-methyl-propyl group is hydroxylated; and M11–M13 (6.9%), which are isomers in which the 2-hydroxyethyl group has been oxidized to an acid and conjugated with glucuronic acid. The remaining metabolites, M1–M4, M6, M7, M9, and M10 each accounted for 3.1% of the radioactivity in urine.

Chapter | 101  Toxicology and Safety Evaluation of Picaridin

Based on the available data on Picaridin, the patterns of metabolism, absorption, distribution, and excretion are similar in humans and rats (Ecker, 1997). The kinetic studies in humans and rats clearly show there is no qualitative difference between Picaridin metabolism in rats and humans. Excretion of KBR 3023 metabolites by humans was substantially faster than by rats, which may be due to the markedly higher percentage of secondary metabolites (glu­ curonides) found in humans compared to the rat. The excretion half-life in humans was 8.2 h compared to 23.3 h in rats at the low dose (Driver et al., 2005).

101.4  Mammalian toxicology The toxicological profile of KBR 3023 is well characterized. All toxicology data were developed using the dermal route of exposure, the most relevant route based on the use pattern of the product. Picaridin is used in products intended solely for dermal application and dermal penetration will certainly be the dominant route of uptake. The rationale of product development using the dermal route of exposure was considered at the suggestion of the U.S. EPA (U.S. Environmental Protection Agency) and in agreement with U.S. EPA, BGA (German authorities) and Health & Welfare Canada. A complete toxicology study required for the registration of an insecticide including acute and subchronic neurotoxicity and metabolism studies was conducted by the dermal route. Additionally 14-day, 5-week, and 14-week dietary feeding studies were conducted to assess any hazard associated with hand-to-mouth transfer from dermal use of Picaridin. In all repeated-dose dermal studies, Picaridin was applied 5 days per week without wiping and covering 10% of the body surface area of the animals, and the administered dose volumes were based on the mean weekly body weight for each dose group and every 3 days for pups in the reproduction study. All application sites were carefully shaved prior to application and in repeated dose studies animals were shaved periodically. Dose application sites were not covered; thus, in order to avoid animal access and subsequent oral ingestion of the test material, all rodent species and rabbits were fitted with Elizabethan collars for the duration of exposures. The collar sizes were adjusted to accommodate the animal’s growth. In the reproduction study, collars were not used during lactation period and pups were fitted with collars after weaning. The highest dermal dose for long-term studies was selected to be 200 mg/kg/day based on the limit of application, to avoid changes in skin integrity (seen at higher doses) for the 18- to 24-month exposure for mouse and rat studies, respectively, and to maximize the systemic dose because the dermal absorption/ kinetic data show that absorption is decreased with increasing dosage, thus lowering overall systemic dose. Given that the dose site was not wiped during the 5-day/week dermal application, the realistic mean dose was at least 50% greater than the target dose. Dermal absorption studies were conducted both

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in the rats and human volunteers to assess the human risk on the absorbed dose analysis associated with the consumer use of the product. Additionally, in vitro dermal penetration studies using human skin were also conducted to compare the dermal absorption of formulations of Picaridin. The toxicology database for Picaridin provided in this chapter was developed by Bayer and LANXESS laboratories to support the registration of commercial products. All toxicology, metabolism, and dermal absorption studies were conducted in compliance with Good Laboratory Practice (GLP) requirements and in accordance with regulatory guidelines of the U.S. EPA (FIFRA), the OECD, and the Japanese MAFF. Human subject studies including dermal absorption, pharmacokinetics, photo­toxicity, and efficacy were conducted with IRB-approved protocols. All toxicology studies were conducted with technical-grade Picaridin having a purity ranging from 96.7 to 98.7%. Published information on the toxicology of Picaridin is limited to publications of the Bayer and LANXESS data.

101.4.1  Acute Toxicity The acute toxicity of Picaridin Technical (KBR 3023) is very low, regardless of the route of administration (Table 101.3). The tests for oral toxicity were conducted on males only, but the studies using other routes of exposure showed no indication for an appreciable gender difference in the toxicity of Picaridin. Picaridin is not irritating to the skin and only slightly irritating to the eye. Picaridin was also negative for skin sensitization in the Buehler patch test on guinea pigs. An additional study with human volunteers demonstrated that Picaridin presented no potential for phototoxic hazard to the test subjects (Lehmann, 1996). Toxicity data for the formulations containing 20% Picaridin (the highest percent formulation on the market) show that Picaridin had low toxicity via all routes of exposure (Table 101.4).

Table 101.3  Summary of the Acute Toxicity Testing with Picaridin (KBR 3023 Technical) Study type

Results

Acute oral – rat

LD50  4743 mg/kg (male) LD50  2236 mg/kg (male)

Acute dermal – rat

LD50  2000 mg/kg LD50   5000 mg/kg (male)

Acute inhalation – rat

LC50  4.364 mg/l

Eye irritation – rabbit

Ocular irritant, slightly

Skin irritation – rabbit

Not a dermal irritant

Dermal sensitization – guinea pig

Not a dermal sensitizer

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Table 101.4  Summary of the Acute Toxicity Testing with 20% Picaridin Formulation Study type Acute oral – rat

Results LD50  5100 mg/kg (male) LD50  5050 mg/kg (male)

Acute dermal – rat

LD50  5040 mg/kg (male) LD50  5020 mg/kg (male)

Acute inhalation – rat

LC50  3.02 mg/l LC50  3.02 mg/l

Eye irritation – rabbit

Ocular irritant, moderate

Skin irritation – rabbit

Not a dermal irritant

Dermal sensitization – guinea pig

Not a dermal sensitizer

101.4.2  Subchronic toxicity 101.4.2.1  Oral In a 5-week toxicity study (Wahle, 2001a), Picaridin was administered in the diet to 10 Sprague-Dawley rats/sex/dose at nominal dose levels of 0, 100, 150, 300, or 1000 mg/kg/ day (actual average daily dose levels of 0, 99, 152, 308, or 1034 mg/kg bw/day in males and 0, 121, 189, 360, or 1141 mg/kg bw/day in females, based on food consumption and body weight data). In a 14-week toxicity study (Wahle 2001b), Picaridin was administered in the diet to 10 Sprague-Dawley rats/ sex/dose at nominal dose levels of 0, 100, 150, 300, or 1000 mg/kg/day (actual average daily dose levels of 0, 100, 149, 301, or 1033 mg/kg bw/day in males and 0, 126, 194, 382, or 1192 mg/kg bw/day in females). In both studies, a decreased body weight gain was noted in both sexes at the highest dose level tested. Effects on the male kidney included an increased relative kidney weight and an increased incidence of protein droplet degenerative nephropathy (possibly due to 2-globulin accumulation). The systemic subchronic no observed adverse effect levels (NOAEL) for the 5- and 14-week oral rat studies were 301 and 308 mg/kg/day, respectively.

101.4.2.2  Dermal In a dermal subchronic 90-day toxicity study (Sheets, 1995), Picaridin was applied to the shaved skin of young adult Sprague-Dawley rats (10–20/sex/dose) at dose levels of 0, 80, 200, 500, or 1000 mg/kg/day for 5 days/week, 5 h/day, for 90 days. Following 90 days of treatment, 10 rats/sex/dose were sacrificed. The remaining 10 rats/sex in

the 0 and 1000 mg/kg/day groups were maintained without treatment for an additional 4 weeks to assess recovery potential. Liver and kidney lesions were seen at dose levels 500 and 1000 mg/kg/day, including increased liver and kidney weight, diffuse liver hypertrophy, necrotic liver cells, slight hyaline degeneration in the kidneys, increased incidence of foci of tubular regeneration, and chronic kidney inflammation. However, there were no clinical pathology anomalies to corroborate these findings. The kidney lesions, seen only in male rats, were associated with 2-globulin and are not relevant to humans. In addition, animals of all treatment groups showed irritative lesions at the application site. These lesions are a common reaction toward repeated application of a variety of materials, including water or medicinal petrolatum, and are not considered a substancerelated adverse effect. After a 4-week recovery period, treated and control groups were similar for all parameters. The systemic subchronic NOAEL for the dermal rat study was 200 mg/kg/day. Based on the irritative lesions seen at dose levels above 200 mg/kg/day, the highest dose tested in all repeated dose studies was selected to be 200 mg/kg/ day to ensure skin integrity was not compromised over the exposure period. In a 13-week toxicity study (Wahle et al., 1999b), technical grade Picaridin was administered dermally to the CD-1 mouse (15 animals/dose/sex) at constant nominal dosages of 0, 80, or 200 mg Picaridin/kg bw/day. All in-life parameters, which included body weight, food consumption, clinical observations, survival, and hematology, were unaffected by dermal exposure to Picaridin. Similarly, postmortem analyses, which included organ weights, gross pathology, and histopathology, were also unchanged following exposure to Picaridin. Thus, the NOAEL was 200 mg/kg/day. Subchronic toxicity studies in the dog were not conducted.

101.4.3  Chronic Toxicity and Oncogenicity 101.4.3.1  Chronic Toxicity in the Dog In a chronic dermal toxicity study (Jones and Hastings, 1995), Picaridin was applied to the clipped skin of beagle dogs (four/sex/dose) at dose levels of 0, 50, 100, or 200 mg/kg/ day, 5 days/week, for one year. Toxicity was not observed in this 1-year dermal chronic toxicity study in the dog. A systemic chronic NOAEL was established in the dog at the highest dose tested (HDT) of 200 mg/kg/day.

101.4.3.2  Combined Chronic Toxicity and Oncogenicity in the Rat The dermal chronic (1-year) toxicity study in rats (Wahle et al., 1999b) was conducted in combination with a 2-year carcinogenicity study. In this study, undiluted Picaridin technical

Chapter | 101  Toxicology and Safety Evaluation of Picaridin

was administered dermally on the dorsal aspect of the trunk to 50 Sprague-Dawley rats/sex/dose at dose levels of 0, 50, 100, or 200 mg/kg/day on 5 consecutive days/week for 24 months. In addition, 10–20 rats/sex/group were killed at 12 months. Picaridin has a very low order of toxicity in the rat following repeated dermal administration. In the dermal chronic toxicity/carcinogenicity study in rats, there were no treatment-related effects produced at any dose level. Adaptive liver changes consisting of cystic degeneration were observed at the HDT of 200 mg/kg/day, but with no corroborating liver weight or clinical pathology anomalies. The rat NOAEL for 1 and 2 years was established at 200 mg/kg/day and no oncogenic potential was observed.

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101.5  Developmental toxicity 101.5.1  Rat A developmental toxicity study in Sprague-Dawley rats was conducted at dose levels of 0, 50, 200, or 400 mg/kg/day. Dams were treated from days 0 through 20 of gestation (Astroff et al., 2000). At the dose site, local dermal reactions (scab and scaling of the skin) were seen on all treated animals. At 400 mg/kg/ day, a slight increase in absolute and relative liver weights was seen and was considered an adaptive response. Treatmentrelated effects on external, visceral or skeletal malformations or external and visceral variations were not seen at any dose level. The maternal and developmental NOAELs were 400 mg/kg/day, the highest dose tested.

101.4.3.3  Oncogenicity in the Mouse In an oncogenicity study (Wahle et al., 1999a), Picaridin was administered dermally on the dorsal aspect of the trunk to 50 CD-1 mice/sex/dose at dose levels of 0, 50, 100, or 200 mg/kg/day on 5 consecutive days/week for 18 months. No toxicity or evidence of an oncogenic potential was observed at any dose level, up to and including the HDT of 200 mg/kg/day. The long-term mouse dermal NOAEL was 200 mg/kg bw/day.

101.4.3.4  Classification for Carcinogenicity Evidence of carcinogenicity was not seen in the rat or mouse oncogenicity studies and Picaridin was negative for genotoxi­ city. On April 22, 1999, EPA Health Effects Division’s Hazard Identification Assessment Review Committee (HIARC) determined that Picaridin was not likely to be a carcinogen by the dermal route, and found no evidence of endocrine disruption.

101.4.4  Genotoxicity Picaridin was non-mutagenic in bacteria and in the mammalian V79 cell line, both with and without metabolic activation (Herbold, 1990, 1991). In an HPRT-forward-mutation study in CHO cells (Brendler, 1991), the Picaridin concentrations tested did not produce sufficient cytotoxicity. In subsequent studies, chromosome aberrations occurred only at cytotoxic concentrations in CHO cells. One study (Gahlmann, 1996) found chromosomal damage in the absence and presence of S9 mix, whereas Gudi and Schadly (1997) detected chromosomal aberrations only when S9 mix was absent. The overall assessment for this endpoint is therefore negative. No unscheduled DNA synthesis was found in rat primary hepato­cytes that were treated with Picaridin (Brendler, 1992). Picaridin did not induce micronucleus formation in murine bone marrow erythroblasts at the maximum tolerated oral dose of 350 mg/kg (Herbold, 1994).

101.5.2  Rabbit Developmental toxicity in the rabbit was conducted at dose levels of 0, 50, 100, or 200 mg/kg bw/day. Animals were treated dermally from days 0 to 28 of gestation (Astroff et al., 2000). A dose-dependent increase was evident for the dermal reactions (slight edema and erythema, cracked skin, squamous skin) to treatment. No toxicity was seen in the rabbit does or fetuses at the maximum dose of 200 mg/ kg/day. NOAEL for maternal and developmental toxicity is 200 mg/kg body weight/day.

101.6  Reproductive toxicity In a two-generation reproduction study (Astroff et al., 1999), Picaridin was administered to the shaved skin of Sprague-Dawley rats (30/sex/dose) at dose levels of 0, 50, 100, or 200 mg/kg/day for 5 days/week. Starting at approximately 8–9 weeks of age, parental (P) animals were administered the test compound dermally for 10 weeks before mating. After the mating period, P males were sacrificed and necropsied. P female dosing continued throughout gestation and lactation. Litters were culled to eight pups/litter on day 4 of lactation and weaned at 21 days of age. P dams were sacrificed and necropsied after weaning. Upon weaning, first-generation (F1) animals were dosed dermally with the same dose of test compound as their dams for 10 weeks before they were mated to produce the second generation (F2). F1 adults and F2 litters were sacrificed and necropsied at weaning. Exposure of all animals to the test material was 5 days/week throughout the study. Dosing was based on the weekly body weight determination, except for F1 pups, where body weight measurements were taken every 3 days from weaning until the start of the premating phase. Elizabethan collars were worn by all P animals for the duration of the study, beginning at least 7 days prior to the initiation of dosing. F1 pups received their collars when

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placed into individual cages, approximately 1 week after weaning. No treatment-related parental systemic toxicity was observed. There were no clinical signs of toxicity or changes in pup weight, viability, or litter sizes noted in the pups at any dose level for the F1 or F2 generations. No treatment-related macroscopic findings in the F1 or F2 pups were observed at any dose level. The NOAEL for parental systemic and developmental toxicity was 200 mg/kg body weight/day.

101.7  Neurotoxicity 101.7.1  Acute In an acute neurotoxicity study (Sheets, 1996a), Picaridin was applied to the shaved skin of young adult Fischer 344 rats (12/sex/dose) for 24 h at dose levels of 0, 200, 600, or 2000 mg/kg. The rats were evaluated for reactions in functional observations and motor activity measurements at 4 h and 7 and 14 days posttreatment.

101.7.2  Subchronic In a subchronic neurotoxicity study (Sheets, 1996b), Picaridin was applied to the shaved skin of young adult Fischer 344 rats (12/sex/dose) at dose levels of 0, 50, 100, or 200 mg/kg/ day, 5 days/week for 13 weeks. The rats were evaluated for reactions in functional observations and motor activity testing at 4 hours and during weeks 4, 8, and 13 of treatment. There was no evidence of systemic toxicity or neuro­ toxicity at the highest doses tested in either the acute (NOAEL  2000 mg/kg) or subchronic (NOAEL  200 mg/ kg/day) neurotoxicity studies.

101.8  Dermal absorption 101.8.1  Rat In a dermal absorption study in rats (Warren and Sturdivant, 1997), animals were treated with [14C] Picaridin at dose levels of 8 mg/kg bw (0.133 mg/cm2), 40 mg/kg bw (0.67 mg/ cm2), or 200 mg/kg bw (3.33 mg/cm2) for 8 h. The absorption patterns were similar between males and females and between all dosage groups. In all treatment groups, the primary route of excretion was via the urine. The majority of radioactivity was excreted within 2 days after application. Total recovery of radiolabel from all groups was high, averaging 95.2% of the applied dose. The data also suggested that the radiotracer did not bind to the application site of the skin. Average absorption (percent of applied dose) was 20.5, 19.1, and 13.8% for males and 27.0, 17.5, and 22.2% for females at respective dose rates of 0.133, 0.67, and 3.33 mg/cm2. The dermal absorption value of 19.1% from the rat study was used for risk assessment as this value corresponded with the dermal dose used in the human study.

Hayes’ Handbook of Pesticide Toxicology

101.8.2  Human Volunteers In a dermal absorption and excretion study in human volunteers, six healthy male volunteers were treated on 4  6 cm2 of the forearm with Picaridin [14C] either as a 15% (w/w) solution in ethanol or as the undiluted technical grade product for 8 h (Selim, 1994). Blood, urine, feces, and tape strippings of the skin were collected. Single dermal applications of 100 l (0.625 mg Picaridin/ cm2) of the 15% ethanol solution or 15 l (0.612 mg Picaridin/cm2) neat Picaridin were well tolerated by the test subjects. Less than 4% of the dermal applied dose of Picaridin, either as a 15% formulation or the neat technical grade substance, was absorbed through the skin during an 8h exposure period and 120 h thereafter. The absorbed part of Picaridin was rapidly excreted via the urine. More than 93% of the absorbed dose was excreted within the first 24 h after exposure had begun. For subjects treated with 15% ethanol solution, Picaridin absorbed was 3.67% of the applied dose, while for treatment with the neat material only 1.66% of the Picaridin was absorbed. There was no evidence that dermal applied Picaridin accumulated in the skin. The dermal absorption studies in humans and rats showed a marked species difference in absorption of neat Picaridin over several days. At equivalent doses (0.625 mg/cm2 in humans; 0.67 mg/cm2 in rats), dermal absorption was considerably greater in rats (19.1% after 7 days) than in humans (1.66% after 5 days) due to greater skin permeability, retention in rat skin (7.6% 16 h after washing), and systemic accumulation. Systemic accumulation is less likely to occur in humans because skin permeability is low, retention in the epidermis of the skin was very low (0.02% 1 h after dosing site wash based on tape stripping, which under normal exposure conditions would be expected to flake off), and excretion was rapid (94% within 24 h). At a comparable dose, humans would be expected to have extremely low systemic Picaridin levels compared to rats, and experience no toxi­ city. Thus, humans are better protected than rats due to low skin permeability and retention, rapid excretion, and a minimum of systemic accumulation. To account for this species difference in risk assessments, a rat:human dermal penetration factor of 11.5 (19.1% ÷ 1.66%) was used in risk assessments.

101.8.3  Human Skin An in vitro study using human skin was conducted with technical Picaridin and 15% ethanolic solution (Rascle, 2008a) at dose levels of 0.63 mg/cm2 of [14C] Picaridin for 8 h. The dermal absorption over 24-h period was 10.66% and 9.812% for Technical and 15% ethanolic solution, respectively. The results indicate that dermal absorption was higher in the in vitro study by a factor of 6.42 and 2.6 for [14C] Picaridin technical and [14C] Picaridin in 15% ethanolic solution respectively.

Chapter | 101  Toxicology and Safety Evaluation of Picaridin

In a similar study conducted with 15% formulation with and without sunscreen (Rascle, 2008b), the results showed that the addition of sunscreen decreased dermal penetration through the human skin. It can be hypothesized that sunscreen acted as a barrier to dermal penetration of Picaridin. In vitro studies conducted by Gu and Chen (2009), using various combinations of Picaridin and an oxybenzone formulation also showed that dermal penetration of Picaridin was suppressed and the degree of suppression was dependent on the concentration and formulation.

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However, the inputs and the NOAELs used in risk calculations by EPA were extremely conservative based on the following: The number of total and consecutive days the repellents are used. l Three applications/day are not warranted based on the long-lasting effectiveness of Picaridin-based repellents. l The NOAEL used for all dermal exposure scenarios was 200 mg/kg/day (nominal concentration) from the repeated dose animal studies. However, the actual dose experienced by the animals was much higher as the dose was repeatedly applied and the dose site was not washed at the end of the exposure day and the dermal absorption in the rat following repeated exposure was higher (50%) compared to the single 8-h exposure (19.1%). l

101.9  Exposure and safety evaluation The risk assessment for formulations containing 20% Picaridin (highest concentration on the market) was conducted on the basis of relative bioavailability of Picaridin derived from the rat and human dermal absorption data (rat/human ratio). The dermal toxicological end point selected by the EPA for risk assessment was 200 mg/kg/day for acute (1-day exposure); short-term exposure (up to 30 days), and intermediateterm exposure (1–6 months) with dermal penetration ratio (rat/human) of 11.5. Use of repellents is generally seasonal, while long-term and chronic exposures are not anticipated. For calculating margin of exposure (MOE) for children’s hand-to-mouth exposure, a NOAEL of 308 mg/kg/day (from oral 5-week study in rats) was used. The consumer use estimates were derived from a study conducted with the most-used repellent DEET (Boomsma et al., 1990). The inputs used in single-dose exposure estimation are provided in Table 101.5. Short- to intermediate-term dose and MOE are provided in Table 101.6. The MOEs for use of the product were 100.

Based on the above observations, the realistic dose experienced by animals was estimated to be 288 mg/kg/day (Driver et al., 2005) and the actual MOEs would be much higher than those provided in Table 101.6. The exposure-risk analyses for the 20% formulations show that sufficient MOEs exist and based on the low skin penetration, rapid excretion of the absorbed material along with high compatibility with skin, Picaridin-based repellents can be safely used by adults and children.

Conclusion The toxicology and safety profile of Picaridin is well characterized. All toxicology data were developed by the dermal route of exposure, the most relevant route based on the use pattern of the product. Picaridin and Picaridin-based products have low acute toxicity by all routes and are not

Table 101.5  Assumptions for Dermal Exposure Assessment and Exposures for Single-Day Exposure Mean body weight (kg)

Mean body Application Rate of application area (25%) surface area (cm2) (cm2)

Adult

70

18,150

4538

1 mg formulation/cm2 skin

Children (3 years)

15

6565

1641

1 mg formulation/cm2 skin

Short-term (single-day) exposures Daily exposure (mg/kg/day) 1 application

2 applications

3 applications

Adult

29.3

58.6

87.9

Children

49.4

98.8

148.2

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Table 101.6  Time-weighted Short-/Intermediate-Term Estimates of Dermal Daily Exposures for Picaridin Number of Applications

Adult (70 kg) Daily exposure (mg/ kg/day)

Child (15 kg) MOE

Daily exposure (mg/kg/day)

MOE

28-day

90-day

28-day

90-day

28-day

90-day

28-day

90-day

1

1.05

0.33

2200

6900

1.76

0.55

1300

4100

2

2.09

0.65

1100

3500

3.53

1.1

640

2100

3

3.14

0.98

720

2300

5.29

1.65

430

1400

Adult 28-day

90-day

Maximum number of consecutive applications

22

69

Maximum number of consecutive days exposed

7

23

Daily exposure (mg/kg/ day)

22.05

28-day

MOE 100

22.77

skin sensitizers. Picaridin showed no neurological or developmental toxicity and there was no evidence of genotoxicity. Chronic dermal dosing in mice, rats, and dogs produced no evidence of adverse toxicity changes and it was not oncogenic in rats and mice. Additional subchronic toxicology studies were conducted by the oral route of exposure. The toxicology profile by the oral route was similar to that of the dermal route and no cumulative effects were evident by oral or dermal routes of exposure. Picaridin is poorly absorbed through the human skin compared to rats and the absorbed dose is rapidly excreted. A conservative risk assessment conducted on the basis of relative bioavailability (rat/human ratio) shows that margins of safety are sufficient and acceptable. Based on favorable toxicology profile, low skin penetration, rapid excretion of the absorbed material along with high compatibility with skin, Picaridin-based repellents can be safely used by adults and children.

References Astroff, A. B., Freshwater, K. J., Young, A. D., Stuart, B. P., Sangha, G. K., and Thyssen, J. H. (1999). The conduct of a two-generation reproductive toxicity study via dermal exposure in the SpragueDawley rat-case study with KBR 3023 (a prospective insect repellent). Reprod. Toxicol. 13, 223–232. Astroff, A. B., Young, A. D., Holzum, B., Sangha, G. K., and Thyssen, J. H. (2000). Conduct and interpretation of a dermal developmental toxicity study with KBR 3023 (a prospective insect repellent) in the Sprague-Dawley rat and Himalayan rabbit. Teratology 61, 222–230. Barnard, D. R., Bernier, U. R., Posey, K. H., and Xue, R. D. (2002). Repellency of IR 3535, KBR 3023, para-menthane-3,8-diol, and

Children 90-day

MOE 100

28-day

90-day

13

41

4

13

22.88

28-day

90-day

MOE 100

MOE 100

22.5

DEET to black salt marsh mosquitoes (Diptera: Culicidae) in the Everglades National Park. J. Med. Entomol. 39(6), 895–899. Boeckh, J., Breer, H., Geier, M., Hoever, F.-P., Kruger, B.-W., Nentwig, G., and Sass, H. (1996). acylated 1,3-aminopropanols as repellents against bloodsucking arthropods. Pestic. Sci. 48, 359–373. Boomsma, J. C., and Parthasarathy, M. (1990). Human use and exposure to insect repellents containing DEET. DPR. Reg. Dot. 50(19 1), 162. Brendler, S. (1991). KBR 3023 – Mutagenicity study for the detection of induced forwardm mutations in the CHO-HGPRT assay in vitro. Bayer AG, Toxicology, Wuppertal, Germany, Report No. 20815, 1991-11-13 (unpublished) Brendler, S. (1992). KBR 3023 – Mutagenicity test on unscheduled DNA synthesis in rat liver primary cell cultures in vitro. Bayer AG, Toxicology, Wuppertal, Germany, Report No. 21314, 1992-04-29 (unpublished). Carpenter, S., Eyres, K., McEndrick, I., Smith, L., Turner, J., Mordue, W., and Mordue, A. J. (2005). Repellent efficiency of BayRepel against Culicoides impunctatus (Diptera:Ceratopogonidae). Parasitol. Res. 10.1007/s00436-005-1298-6. Carroll, S. P. (2008a). Efficacy Test of KBR 3023 (Picaridin; Picaridin)based personal insect repellents (20% Cream and 20% Spray) with mosquitoes under field conditions. Carroll-Loye Biological Research, Davis, California, Laboratory Project I.D. LNX-001. Carroll, S. P. (2008b). Efficacy test of picaridin-based personal repellents with mosquitoes under field conditions. Carroll-Loye Biological Research, Davis, California, Laboratory Project I.D. SPC-001. Carroll, S. P. (2008c). Efficacy test of Picaridin-based personal tick repellents. Carroll-Loye Biological Research, Davis, California, Laboratory Project I.D. SPC-002. Costantini, C., and Ilboudo-Sanogo, E. (2001). WHOPES evaluation of insect repellent KBR 3023 in Burkina Faso. Final Report for WHO Project V2.181.276. Diesing, L. (1991). Study for skin-sensitizing effect on guinea pigs – components of KBR 3023: KBR 4230, KBR 4223, 2-(2-hydroxyethyl)-piperidine,

Chapter | 101  Toxicology and Safety Evaluation of Picaridin

chloroformic acid secondary butyl ester, anhydride of chloroformic acid secondary butyl ester. Bayer AG, Dept. of Toxicology, Wuppertal, Germany, Report No.: PH-20046; Bayer AG Tox Study Nos. T3037030, T4037031, T8037224, T9037225, and T0037226; 1991-03-07 (unpublished). Dreist, M. (1991). Study for skin-sensitising effect on guinea pigs. Bayer AG, Dept. of Toxicology, Wuppertal, Germany, Report No.: 20623, 1991-09-05 (unpublished), amended by Report No.: 20623A; 1999-08-30. Driver, J. H., Ross, J. H., and Sangha, G. K. (2005) KBR 3023-based insect repellents: Probabilistic exposure & risk analysis – 20% formulation. LANXESS Corporation, 111 Park West Drive, Pittsburgh, PA, USA, Project Identification: 01-LX-05, 2005-09-23 (unpublished). Ecker, W. (1997). [Hydroxyethyl-1-14C] KBR 3023: Human volunteer metabolism study after dermal application. Bayer AG Institute for Metabolism Research and Residue Analysis, Leverkusen, Germany, Report No. PF 4187, 1997-01-07 (unpublished). Ecker, W., and Weber, H. (1997). [Hydroxyethyl-1-14C] KBR 3023: Rat metabolism study after intravenous injection and after dermal application. Bayer AG Institute for Metabolism Research and Residue Analysis, Leverkusen, Germany, Report No. PF 4178, 1997-02-27 (unpublished). Frances, S. P., Dung, N. V., Beebe, N. W., and Debboun, M. (2002). Field evaluation of repellent formulations against daytime and nighttime biting mosquitoes in a tropical forest in Northern Australia. J. Med. Entomol. 39(3), 541–544. Frances, S. P., Waterson, D. GE., Beebe, N. W., and Cooper, R. D. (2004). Field evaluation of repellent formulation containing DEET and Picaridin against mosquitoes in Northern Territory, Australia. J. Med. Entomol. 41(3), 414–417. Gahlmann, R. (1996). KBR 3023 – In vitro mammalian chromosome aberration test with chinese hamster ovary (CHO) cells. Bayer AG, Toxicology, Wuppertal, Germany, Report No. 25019, 1996-04-30 (unpublished). Gu, X., and Chen, T. (2009). In vitro permeation characterization of repellent Picaridin and sunscreen oxybenzone. Pharm. Devel. Tech. 14, 485–491. Gudi, R., and Schadly, E.H. (1997). Chromosome aberrations in Chinese hamster ovary (CHO) cells. Microbiological Associates, Inc., Rockville, MD, USA, Report No. 107777, 1997-08-04 (unpublished). Herbold, A. (1990). KBR 3023 – Salmonella/microsome test. Bayer AG, Toxicology, Wuppertal, Germany, Report No. 18917, 1990-03-16 (unpublished). Herbold, B. (1991). KBR 3023 – V79/HPRT-test in vitro for the detection of induced forward mutations. Bayer AG, Toxicology, Wuppertal, Germany, Report No. 29220, 1999-10-19 (unpublished). Herbold, B. (1994). KBR 3023 – Micronucleus test on the mouse. Bayer AG, Toxicology, Wuppertal, Germany, Report No. 23291, 1994-08-29 (unpublished). Herbold, A. (1999). KBR 8180 By-product of KBR 3023 Salmonella/ microsome test plate incorporation and preincubation method. Bayer AG, Toxicology, Wuppertal, Germany, Report No. PH-28459, 1999-02-08 (unpublished). Jones, R. D., and Hastings, T.F. (1995). Technical grade KBR 3023: A chronic percutaneous toxicity study in the beagle dog. Bayer Corporation, Agriculture Division, Toxicology, Stilwell, KS, USA, Report No. 107155, 1995-12-01 (unpublished). Krötlinger, F. (1988). KBR 3023: Investigations of the acute dermal toxicity to rats. Bayer AG, Dept. of Toxicology, Wuppertal, Germany, Report No. PH-16995, Bayer AG Study No. T7029600, 1988-05-25 (unpublished). Krötlinger, F. (1990a). KBR 3023 study for acute oral toxicity to rats. Bayer AG, Dept. of Toxicology, Wuppertal, Germany, Report No. 19295, 1990-07-20, (unpublished).

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Krötlinger, F. (1990b). KBR 8180 (N-methyl-O-ester) (byproduct of KBR 3023) study for acute oral toxicity to rats. Bayer AG, Dept. of Toxicology, Wuppertal, Germany, Report No. 28483, 1999-02-16, (unpublished). Lehmann, P. (1996). Controlled intra-individual comparative study of the phototoxicity of the repellent KBR 3023. Heinrich-Heine-University, Skin Clinic, Düsseldorf, Germany, Report No. 107792, 1996-06-31 (unpublished). Luepkes, K. H. (2000). Mosquito repellent effects of various formulations based on Bayrepel/KBR 3023 against yellow fever mosquito Aedes aegypti. Unpublished Lanxess report. Muelhofer, (1993). LHS-Labor for Hygiene and Sicherheit (Laboratory for Hygiene and safety, Austria). Unpublished LANXESS report. Mordue (Luntz), A. J. (1999). Field evaluation of the repellent KBR 3023 against the Scottish biting midge, Culicoides impunctatus. Department of Zoology, University of Aberdeen, Tillydrone Avenue, Aberdeen AB24 2TZ. WHOPES Report. Nentwig, G. (1997). Efficacy of a formulation with 12% KBR 3023 in comparison to a formulation with 17% Deet on human arms against three different blood sucking diptera. Bayer Report No. 018032. Nentwig, G. (1998). Efficacy of KBR 3023 in comparison to Deet on human arms against the cat flea, Ctenocephalides felis. ID No. 018580 (BG Animal Health). Perrotey, S., Madulo Leblond, G., and Pesson, B. (2002). Laboratory testing of the insect repellent KBR 3023 against Phlebotomus duboscqi (Diptera: Psychodidae). Parasitol. Res. 88(7), 712–713 DOI 10.1007/s00436-002-0635-2. Peterson, C., and Coats, J. (2001). Insect repellents – past, present and future. Pestic. Outlook 154–158. Rascle, J. B. (2008a). Cutter insect repellent SS and cutter insect repellent formula A formulations comparative in vitro dermal absorption study using human skin. Bayer CropScience, rue Dostoievski, Sophia Anipolis Cedex, France. Report No.: SA 07027, 2008-01-31, (unpublished). Rascle, J. B. (2008b). KBR 3023 technical and KBR 3023 15% in ethanol comparative in vitro dermal absorption study using human and rat skin. Bayer CropScience, rue Dostoievski, Sophia Anipolis Cedex, France. Report No.: SA 07026, 2008-01-11, (unpublished). Selim, S. (1994). A single dose open label study to investigate the absorption and excretion of a 14C-labelled insect repellent (KBR 3023) from two different formulations after dermal application to healthy volunteers. Clinical part: Pharma Bio-Research Laboratories, Zuidlaren, The Netherlands. Analytical part: Biological Test Center, Irvine, CA, USA, Report-No. P1092004, 1994-06-20 (unpublished). Sheets, L. P. (1995). A repeated dose 90-day dermal toxicity study with technical grade KBR 3023 in rats. Bayer Corporation, Agriculture Division, Toxicology, Stilwell, KS, USA, Study No. 90-122-HC, 1995-11-01 (unpublished). Sheets, L. P. (1996a). An acute dermal neurotoxicity screening study with technical grade KBR 3023 in Fischer 344 rats. Bayer Corporation, Agriculture Division, Toxicology, Stilwell, KS, USA, Report No. 107467, 1996-10-14 (unpublished). Sheets, L. P. (1996b). Subchronic dermal neurotoxicity screening study with technical grade KBR 3023 in Fischer 344 rats. Bayer Corporation, Agriculture Division, Toxicology, Stilwell, KS, USA, Report No. 107466, 1996-10-09 (unpublished). Sheets, L. P., and Philipps, S. D. (1991). Acute dermal toxicity study with technical grade KBR3023 in rats. Mobay Corporation, Corporate Toxicology Department, Stilwell, KS, USA, Report No. 101284, Study No. 90-022-GD, 1991-08-27 (unpublished).

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Sixl, W. (1993). Efficacy study with three preparations called Autan 611, Autan 617 and Autan 619. Geomedizinisce Forschungsstelle (Geomedical Research Office); Parasitologische Untersuchungsund Beratumgsstelle (Parasitological Examination and Consulting Agency); Hygiene Institute der Universitat (University Hygiene Institute), Austria. Unpublished LANXESS Report. Todd, R. (1997). Evaluation of a personal repellent against deer ticks ICR – Insect Control and Research, Inc. Unpublished LANXESS Report. Wahle, B. S. (2001a). Technical grade KBR 3023: A subchronic toxi­ city testing study in the rat (5-week interval). Bayer Corporation, Agriculture Division, Toxicology, Stilwell, KS, USA, Report No. 110222, 2001-04-05 (unpublished). Wahle, B. S. (2001b). Technical grade KBR 3023: A subchronic toxi­ city testing study in the rat (14-week interval). Bayer Corporation, Agriculture Division, Toxicology, Stilwell, KS, USA, Report No. 110223, 2001-04-05 (unpublished). Wahle, B. S., Sangha, G. K., Elcock, L. E., Sheets, L. P., and Christenson, W. R. (1999a). Carcinogenicity testing in the CD-1 mouse of a prospective insect repellant (KBR 3023) using the dermal route of exposure. Toxicology 142, 29–39.

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Wahle, B. S., Sangha, G. K., Lake, S. G. L. E., Sheets, L. P., Croutch, C., and Christenson, W. R. (1999b). Chronic toxicity and carcinogeni­ city testing in the Sprague-Dawley rat of a prospective insect repellant (KBR 3023) using the dermal route of exposure. Toxicology 142, 41–56. Warren, D. L., and Sturdivant, D. W. (1997). Dermal absorption of technical KBR 3023. Bayer Corp., Agricultural Division, Toxicology, Stilwell, KS, USA, Report No.: 107488, 1997-01-28, (unpublished). World Health Organization. (2000). Review of: IR3535; KBR3023; (RS)-Methoprene 20% EC, Pyriproxyfen 0.5% GR; and Lambdacyhalothrxn 2.5% CS. Report of the Fourth WHOPES Working Group meeting, Geneva, 12-4,5-2000. Yap, H. H., Jahnaqir, k., Chong, A. S., Adanan, C. R., Chong, N. L., Malik, Y. A., and Rohaizat, B. (1998). Field efficacy of a new repellent, KBR 3023, against Aedes albopictus (SKUSE) and Culex quinquefasciatus in a tropical environment. J. Vector Ecol. 23(1), 62–68. Yap, H. H., Jahangir, K., and Aairi, J. (2000). Field efficacy of four insect repellent products against vector mosquitoes in a tropical environment. J. Mosq. Control Assoc. 16(3), 241–244.

Chapter 102

Chlorantraniliprole: An Insecticide of the Anthranilic Diamide Class Karin S. Bentley, Joan L. Fletcher and Michael D. Woodward DuPont Crop Protection, E. I. du Pont de Nemours and Company, Newark, Delaware

102.1  Introduction Chlorantraniliprole is a new insecticide belonging to the anthranilic diamide class of chemistry and is intended for the control of Lepidopteran, Coleopteran, and some Dipteran pests in commercial agriculture on both perennial and annual crops. This product is suitable for field and protected crops and will also be used in some nonagricultural applications, specifically in turf and ornamentals in some countries. Chlorantraniliprole is a residual insecticide active on larvae as well as some adults by both ingestion and contact routes of entry. Ingestion is the most effective method of entry and typically requires a lower dose for response. Chlorantraniliprole provides excellent plant protection since affected insects cease feeding almost immediately after contact with the product. The mode of action of Chlorantraniliprole is currently only shared with one other commercial insecticide active substance, flubendiamide. There is no cross resistance to other insecticides. The regulatory review of this new active ingredient was conducted in 2007 as a global joint assessment under the auspices of the Organization for Economic Co-Operation and Development (OECD). The evaluation was conducted by technical experts from Australia, Canada, Ireland, the United States and the United Kingdom, and peer reviewed by several additional European authorities. Chlorantraniliprole was subsequently reviewed for the first time in 2008 by the Joint FAO/WHO Meeting on Pesticide Residues for the purposes of establishing maximum residue levels.

102.2  Identity, properties, and uses 102.2.1  Chemical Name Chlorantraniliprole (Figure 102.1) is 3-bromo-4-chloro1-(3-chloro-2-pyridyl)-2-methyl-6-(methylcarbamoyl) Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

N Cl

O N O N

Br N N Cl

Figure 102.1  Structure of Chlorantraniliprole.

pyrazole-5-carboxanilide (according to IUPAC) or 3-bromoN-[4-chloro-2-methyl-6-[(methylamino)carbonyl] phenyl]-1-(3-chloro-2-pyridinyl)-1H-pyrazole-5-carboxamide (according to Chemical Abstracts).

102.2.2  Synonyms The common name Chlorantraniliprole (ISO approved) is in general use. Trade names include Rynaxypyr for the technical material, and Altacor, Coragen, Prevathon, Dermacor X-100, and Ferterra for the formulations. The CAS registry number is 500008-45-7, and the DuPont development code is DPX-E2Y45.

102.2.3  Physical and Chemical Properties Chlorantraniliprole has the molecular formula of C18H14BrCl2N5O2 and a molecular weight of 483.15. It is a fine, crystalline off-white solid with a melting point of 200–202°C and has a very low vapor pressure. It is relatively insoluble in most solvents and has low solubility in water (0.88 mg/l at pH 7). The octanol/water partition 2231

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coefficient (LogPow) of Chlorantraniliprole is 2.86 at pH 7; therefore, bioaccumulation is unlikely. Chlorantraniliprole is stable to hydrolysis under neutral and acid conditions but degrades under alkaline conditions (half-life at pH 9 is 10 days). The photolytic half-life of Chlorantraniliprole in sterile aqueous buffer solution (pH 7.0) under continuous irradiation was 0.37 days. Chlorantraniliprole is not flammable, autoflammable, oxidizing, or explosive.

Findings indicate that Chlorantraniliprole exhibits excellent differential selectivity for insect ryanodine receptors over mammalian ryanodine receptors. This selectivity is likely a major contributing factor to the mammalian safety observed with Chlorantraniliprole.

102.2.4  History, Formulations, and Uses

102.4.1  Toxicokinetics in Rats

DuPont introduced Chlorantraniliprole in 2007. Formulations include a WG (water-dispersible granular) formulation (Alt­ acor), two SC (suspension concentrate) formulations with different concentrations (Coragen and Prevathon), a seed treatment formulation (Dermacor X-100), and a low-strength granular formulation (G) named Ferterra. Formulations containing Chlorantraniliprole are marketed in North and South America, Europe, Asia, and Australia.

102.4.1.1  Oral Administration

102.3  Biological mode of action Chlorantraniliprole controls insects through unregulated activation of ryanodine receptor channels leading to internal calcium store depletion that impairs regulation of muscle contraction. Insects exposed to Chlorantraniliprole exhibit general lethargy and muscle paralysis followed ultimately by death. It has been shown that Chlorantraniliprole and other compounds belonging to the anthranilic diamide class of insecticides are highly potent activators of insect ryanodine receptors. These compounds have a novel mode of action for synthetic insecticides (Cordova et al., 2006; Gutteridge et al., World Patent Application WO2004027042, 2004; Lahm et al., 2005, 2007; Sattelle et al., 2008). In physio­logical studies conducted using neurons isolated from the American cockroach, Periplaneta americana, Chlorantraniliprole stimulates release of calcium stores via selective ryanodine receptor activation. The process of insect muscle contraction involves modulation of two distinct calcium channels, voltage-gated calcium channels that regulate external calcium entry and ryanodine receptor channels that regulate release of internal calcium stores from the sarcoplasmic reticulum. Under normal physiological conditions, input from the nervous system activates voltage-gated calcium channels leading to a rise in intracellular calcium. This rise triggers activation of ryanodine receptor channels located on the sarcoplasmic reticulum, resulting in release of stored pools of internal calcium and initiating a muscle contraction. Following such a muscle contraction, the ryanodine receptor channels close allowing calcium stores to refill in preparation for the next stimulated contraction.

102.4  Absorption, distribution, metabolism, and excretion

The absorption and elimination of Chlorantraniliprole were evaluated in male and female rats following low (10 mg/kg bw) or high (200 mg/kg bw) single oral doses. Additional animals of both sexes received 14 consecutive doses (10 mg/kg bw/day) followed by a 7-day depuration period. Two radiolabeled forms of Chlorantraniliprole were used to evaluate the metabolic fate of the active substance in animals. The 14C-radiolabels were located on either the pyrazole or benzamide carbonyl carbon. Essentially all metabolites retained the complete structure of the parent molecule, confirming that the positions of the radiolabels were metabolically stable. The absorption of 14C-Chlorantraniliprole was rapid with peak plasma concentrations occurring at 5–12 h after lowor high-dose administration. Absorption at the low-dose (10 mg/kg bw) was determined to be 73–85% compared with 12–13% at the high dose (200 mg/kg bw) based on the sum in urine, bile, and carcass (excluding gastrointestinal contents). The plasma elimination half-lives were shorter in males (38–43 h) than in females (78–82 h). Tissue distribution of the absorbed dose was extensive, and tissue:plasma ratios by terminal sacrifice were substantially less than 1, indicating low potential for accumulation. The tissue residues were higher in female than male rats consistent with female rats having a longer elimination half-life and higher area under the plasma concentration vs. time curve (AUC). Excretion was substantially complete by 48–72 h after dosing. Fecal excretion was the primary route of elimination (62–64% for low dose; 91–92% for high dose) followed by the urine (24–29% for low dose, 4–5% for high dose) with no significant excretion occurring by exhalation. Steady-state kinetic behavior after 14 days of dose administration was apparent in the male rats and was near steady state in female rats. An evaluation of tissue residues in 21 different tissues produced concentration and percent of dose profiles that were similar to those observed in the single-dose study. Tissue:plasma concentration ratios were less than 1, confirming minimal potential for accumulation. The overall pattern of urinary (12–17%) and fecal (73–82%) elimination for multiple dosing (10 mg/kg bw/day  14 days)

Chapter | 102  Chlorantraniliprole: An Insecticide of the Anthranilic Diamide Class

generally resided between the patterns observed for single low- (10 mg/kg bw) and high- (200 mg/kg bw) dose administration.

102.4.1.2  Dermal Administration The dermal absorption of Chlorantraniliprole formulated both as a WG formulation and as a SC formulation was determined in vitro using rat and human skin and in vivo in rats. Testing was performed with the undiluted formulation concentrates and a 0.75 g/l aqueous dilution designed to resemble field use concentrations. During the 24-h monitoring period in the in vitro dermal absorption study, Chlorantraniliprole did not penetrate rat or human skin to an appreciable degree with either formulation. The total absorbable dose in most cases was based on the portion of the dose retained in the skin only. For the concentrate, a greater portion of the dose was retained in rat skin in comparison with human skin. For the aqueous dilution, there was little difference in the total absorbable dose between rat and human skin. In vivo, a low percentage of the applied dose of Chlorantraniliprole (both formulations) was absorbed following a 6-h dermal application to rats. At the final 504-h sampling point, 0.3 and 3% of the applied dose for the SC formulation and 1 and 2% of the applied dose for the WG formulation was absorbed (including the quantity of radioactivity retained in tape-stripped skin) from the undiluted concentrate and the aqueous dilution, respectively. As there was essentially no penetration of Chlorantraniliprole observed in the in vitro dermal absorption study, dermal absorption values for risk assessment were based on the in vivo rat data alone.

102.4.2  Metabolic Pathways in Rats The metabolic fate of Chlorantraniliprole in rats was determined following investigation of residues in urine, feces, and bile samples obtained from the toxicokinetic studies noted above. Metabolism of the absorbed dose was extensive and involved sex differences primarily in initial methylphenyl and N-methyl carbon hydroxylation. Further metabolism of the hydroxylated metabolites included N-demethylation, nitrogen-to-carbon cyclization with loss of a water molecule, oxidation of alcohols to carboxylic acids, amide bridge cleavage, amine hydrolysis, and O-glucuronidation (Figure 102.2). Most of the administered dose (88–97%) was eliminated in the excreta. The profile of metabolites at the low and high dose was similar except that a much greater proportion of the administered dose was recovered as unmetabolized parent Chlorantraniliprole in the feces at the high dose. The profile of metabolites in urine and feces during multiple-dose administration was similar to that observed for the single-dose study.

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102.4.3  Plasma Concentrations of Parent and Metabolites in Feeding Studies In addition to the rat metabolism studies conducted with radiolabeled Chlorantraniliprole, plasma samples were screened during the 90-day studies in rats, mice, and dogs to evaluate dietary exposure and internal dose. Additionally, the plasma pharmacokinetics of parent Chlorantraniliprole was evaluated in male rats after a 14-day oral gavage study. The results clearly demonstrated systemic uptake and metabolism of Chlorantraniliprole during oral administration. The results also suggested possible species differences in the primary metabolites formed by rats, mice, and dogs. The concentration of Chlorantraniliprole in plasma was dog  rat  mouse. The primary methylphenyl ring hydroxylated metabolite, IN-HXH44, was quantified only in dog plasma; whereas the other primary N-methyl hydroxylated metabolite, IN-H2H20, was quantifiable only in rat plasma. IN-GAZ70, a cyclization product of IN-H2H20 with loss of a water molecule or an N-demethylation product of IN-EQW78, was quantified in mouse and rat, but not dog plasma. Mouse plasma contained substantially more IN-GAZ70 than rat plasma. In all three species, the relatively constant analyte concentrations at the higher dose levels suggested decreased relative absorption. The slight decrease in the plasma Chlorantraniliprole concentrations with increasing dose in the 14-day oral gavage study also provided evidence for decreased absorption. A significant sex difference was observed in rats, with female rats showing higher concentrations of parent, IN-H2H20, and IN-GAZ70 than male rats. Sex differences were not apparent for the dog or mouse. In general, the results in rats for the 90-day and 14-day studies were consistent with the plasma concentrations of 14C residues observed in the metabolism studies, with the decreased relative absorption observed at higher-dose levels, and with the proposed meta­ bolic pathway from the single and multiple oral gavage studies with 14C-Chlorantraniliprole (Figure 102.2).

102.5  Toxicity to laboratory animals 102.5.1  Acute Toxicity The acute toxicity of Chlorantraniliprole is summarized in Table 102.1. Chlorantraniliprole has no significant acute toxicity via the oral, dermal, and inhalation routes of expos­ ure. It is not an eye or skin irritant and does not cause skin sensitization. Based on these results, no classification of Chlorantraniliprole for acute toxicity is required according to the provisions of EU Directive 67/548/EEC. According to the criteria of the U.S. EPA, Chlorantraniliprole is classified in Toxicity Category IV.

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2234

Cl

Cl

N

N

Br

N N

N

Cl

NH

N

O

N N

N

N

DPX-E2Y45

HO

Br

IN-H2H20 & H2H20-O-G

Cl

O

HO

Br

Cl

IN-K9T00 & K9T00-O-G

IN-HXH40 & HXH40-O-G

Br

NH N

Br N N

HO

Br N N

N

Cl

Cl IN-HXH44 & HXH44-O-G

N N

N

O

N

HO

O

Cl

N

Br

N N

O

Cl

N

NH

N

N N

O

O

HO

NH2 Cl

NH

O N

HN Cl

IN-GKQ52

NH

N N

Br

Cl

O

Cl

Cl

N N

N

OH

HN Cl

O

Br

Br

IN-F9N04

NH

O

O

Cl

O

NH

N N

N

OH

HN Cl

NH

O

IN-GAZ70

HN Cl

OH

NH

Cl

Cl

Cl

O

Br

N N

N

IN-EQW78

O

NH2

O

O

Cl

IN-K3X21

IN-K7H29 & K7H29-O-G

HN Cl

O

O

Cl

N

NH

HO O O N

N

Br

N N

O Cl

N N

O N

HO

Cl

Cl IN-KAA24

N

Br HO

O N

N N

Br

Cl IN-LQX30 & LQX30-O-G

IN-LEM10 HN

OH

Cl

O N

NH

N N Cl

IN-DBC80

Br

O

+

NH2 HO

O

IN-L8F56

Figure 102.2  Proposed metabolic pathway of chlorantraniliprole (DPX-E2Y45) in the rat. Note that the identification of O-glucuronic acid conjugates is indicated by the addition of “-O-G” to the applicable metabolite IN-code names.

Chapter | 102  Chlorantraniliprole: An Insecticide of the Anthranilic Diamide Class

2235

Table 102.1  Summary of Acute Toxicity Studies with Chlorantraniliprole Type of study

Species

Results

Acute oral LD50

Rat

5000 mg/kg bw

Acute dermal LD50

Rat

5000 mg/kg bw

Acute inhalation LC50 (4 h)

Rat

5.1 mg/l

Skin irritation

Rabbit

Not irritating

Eye irritation

Rabbit

Slight irritation clearing within 72 h

Skin sensitization (Maximization test)

Guinea pig

Not sensitizing

Skin sensitization (Local lymph node assay)

Mouse

Not sensitizing

102.5.2  Subchronic Toxicity

102.5.2.2  Rat, 28-Day Feeding

The following short-term oral toxicity studies were conducted with Chlorantraniliprole: 14-day oral gavage study in rats, 28-day feeding studies in rats, mice and dogs, 28-day capsule administration study in dogs, 90-day feeding studies in rats, mice and dogs, and a 1-year feeding study in dogs. A 28-day percutaneous toxicity study was also performed with Chlorantraniliprole in rats.

In a 28-day feeding study, Chlorantraniliprole was administered to male and female Crl:CD(SD)IGS BR rats (five animals/sex/concentration) at 0, 300, 1500, or 8000 ppm (equivalent to 0, 21/24, 106/128, or 584/675 mg/kg bw/day males/females). (Due to limited availability of material at the time the study was conducted, 8000 ppm was the highest dose tested.) No treatment-related adverse effects on survival, clinical observations, body weight, food consumption, or clinical and gross pathology were observed in males or females over the 28-day interval. A slight increase in liver weight in the 1500- and 8000-ppm females was considered test substance-related but nonadverse. This finding was accompanied by a minimal increase in hep­ atocellular hypertrophy in the 8000-ppm females and was attributed to enzyme induction. Increases in hepatic UDPglucuronyl transferase content occurred in the 1500- and 8000-ppm males and females.

102.5.2.1  Rat, 14-Day Oral Gavage A 2-week oral dosing study was conducted with Chlorantra­ niliprole at dose levels of 0, 25, 100, and 1000 mg/kg bw/day with groups of five Crl:CD(SD)IGS BR rats/sex dosed for 14 consecutive days. There were no test substance-related effects on in-life parameters at any level tested. All animals survived to the end of the study, and the data for clinical observations, body weights, and weight gains were compar­ able across all groups on study. No test substance-related effects were evident in the data collected at or after study termination including clinical pathology (hematology, coagulation, clinical chemistry, and urinalysis), organ weights, and gross and microscopic anatomic pathology. In a toxicokin­ etic assessment, the plasma AUC was not proportional with the dose of Chlorantraniliprole, indicating decreased relative absorption at higher doses. Chlorantraniliprole did not alter hepatic -oxidation activity or cytochrome P450 enzymes in male rats, or total hepatic cytochrome P450 content in female rats. In females, Chlorantraniliprole was a weak inducer of cytochrome P450 isozyme 3A. This enzyme induction was considered to be test substancerelated but not adverse and consistent with a pharmacologic response to increased metabolism. No increases in the frequency of micronucleated polychromatic erythrocytes (PCEs) were observed at any dose level in bone marrow of male and female rats taken at the termination of the dosing period.

102.5.2.3  Rat, 28-Day Percutaneous In a repeated dose dermal study, Chlorantraniliprole was applied to the shaved, intact dorsal skin of male and female Crl:CD(SD)IGS BR rats (10 animals/sex/dose). The test substance was applied for 29 daily (consecutive) applications. The rats were exposed to the test substance for 6 h per day. Exposure doses were 0, 100, 300, or 1000 mg/kg bw/day. No test substance-related effects were observed on survival, clinical observations, body weights, or food consumption. No adverse test substance-related effects were observed on organ weights and any clinical pathology, gross, or microscopic pathology endpoint. Test substancerelated reductions in mean body weight gain and food efficiency were observed over the 28-day interval in males and females at the highest dose, 1000 mg/kg bw/day. However, there was no effect of treatment on overall body weight. A minimal increase in microvesiculation in the zona

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fasciculata region of the adrenal cortex was observed in some males at 1000 mg/kg bw/day. This finding was considered test substance-related but not adverse as the adr­enal morphology was within the range that may be observed in control rats and was not associated with any indication of cytotoxicity or other evidence of structural or functional impairment of the adrenal gland (see later discussion regarding effects on adrenal function).

102.5.2.4  Mouse, 28-Day Feeding In a 28-day feeding study, Chlorantraniliprole was administered to male and female Crl:CD1(ICR) mice (10 animals/ sex/concentration) at 0, 300, 1000, 3000, or 7000 ppm (equivalent to 0, 52/64, 182/206, 538/658, or 1443/1524 mg/ kg bw/day males/females). Test substance-related reductions in mean body weight gain and food efficiency were observed over the 28-day interval in males at the highest dose. These findings were considered spurious as similar changes were not observed in feeding studies in mice of longer durations. A slight increase in mean liver weight parameters observed in females fed 3000 or 7000 ppm and a mild increase in cytochrome P450 content observed in males and females fed 3000 or 7000 ppm were considered nonadverse pharmacological responses attributed to enzyme induction.

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102.5.2.7  Rat, 90-Day Feeding In a 90-day feeding study, Chlorantraniliprole was administered to male and female Crl:CD(SD)IGS BR rats (10 animals/sex/concentration) at 0, 600, 2000, 6000, or 20,000 ppm (equivalent to 0, 37/47, 120/157, 359/460, or 1188/1526 mg/kg bw/day males/females). No test substance-related effects were observed on survival, body weight and nutritional parameters, or clinical and ophthalmological observations. No adverse test substance-related effects were observed on organ weights or on any clinical, gross or microscopic pathology endpoint. A slight increase in liver weight in 20,000-ppm females and a reduction in bilirubin in females exposed to 2000 ppm and above were considered test substance-related but not adverse. These findings were not associated with any liver histopathology and were attributed to enzyme induction. A minimal to mild increase in microvesiculation in the zona fasciculata region of the adrenal cortex was observed in some males at 20,000 ppm. This finding was considered test substancerelated but not adverse as the adrenal morphology was within the range that may be observed in control rats and was not associated with any indication of cytotoxicity or other evidence of structural or functional impairment of the adrenal gland.

102.5.2.8  Mouse, 90-Day Feeding 102.5.2.5  Dog, 28-Day Oral Capsule In a 28-day oral study, Chlorantraniliprole was administered to male and female beagle dogs (two animals/sex/dose) in gelatin capsules at dosages of 300 or 1000 mg/kg bw/day. An additional group (two dogs/sex) received empty gelatin capsules. No treatment-related effects on survival, body weight or nutritional parameters, clinical or neurobehavioral findings, ophthalmology, clinical pathology, or anatomic pathology were attributed to exposure to Chlorantraniliprole. Oral exposure of male and female dogs to 1000 mg/kg bw/day induced increases in hepatic cytochrome P450 enzyme content. Increases were also observed in isozymes 1A1 and 2B1/2 at 300 and 1000 mg/kg bw/day.

102.5.2.6  Dog, 28-Day Feeding In a 28-day feeding study, Chlorantraniliprole was administered in the diet to male and female beagle dogs (two animals/sex/concentration) at escalating concentrations of 1000 (week 1), 5000 (week 2), and 10,000 (weeks 3–4) ppm for one group and 30000 (weeks 1–2) and 40,000 (weeks 3–4) ppm for the other group. No adverse test substance-related effects were observed throughout the study on survival, clinical or neurobehavioral findings, body weight or weight gain, food consumption, or food efficiency. No adverse test substance-related effects on gross macroscopic pathology were noted in any dog.

In a 90-day feeding study, Chlorantraniliprole was administered to male and female Crl:CD-1(ICR)BR mice (15 animals/sex/concentration) at 0, 200, 700, 2000, or 7000 ppm (equivalent to 0, 33/41, 115/158, 345/422, or 1135/1539 mg/ kg bw/day males/females). No test substance-related effects were observed on survival, body weight and nutritional parameters, or clinical and ophthalmological observations. No adverse test substance-related effects were observed on organ weights or on any clinical, gross or microscopic pathology endpoint. A slight increase in liver weight in 7000-ppm males and females was considered test substance-related but not adverse. The increased liver weights were not associated with any liver histopathology and were attributed to enzyme induction.

102.5.2.9  Dog, 90-Day Feeding In a 90-day feeding study, Chlorantraniliprole was administered to male and female beagle dogs (four animals/sex/ concentration) at 0, 1000, 4000, 10,000, or 40,000 ppm (equivalent to 0, 32/37, 119/133, 303/318, or 1163/1220 mg/ kg bw/day males/females). No test substance-related effects were observed over the 90-day interval on mean body weight gain and nutritional parameters or on clinical, neurobehavioral, and ophthalmological signs in dogs. No test substance-related effects on clinical, gross or microscopic pathology were observed in dogs exposed to any

Chapter | 102  Chlorantraniliprole: An Insecticide of the Anthranilic Diamide Class

dietary concentration. A mild increase in liver weight was observed in male dogs and was statistically significant at 40,000 ppm. This finding was not associated with any liver histopathology and was attributed to enzyme induction.

102.5.2.10  Dog, 1-Year Feeding In a 1-year feeding study, Chlorantraniliprole was administered to male and female beagle dogs (four animals/sex/ concentration) at 0, 1000, 4000, 10,000, or 40,000 ppm (equivalent to 0, 32/34, 112/113, 317/278, or 1164/1233 mg/ kg bw/day males/females). No test substance-related effects were observed on survival, clinical and neurobehavioral signs, ophthalmology, body weight and nutritional parameters, clinical pathology, or gross or microscopic pathology. Test substance-related increases in liver weight were observed in 40,000-ppm male and female dogs, but were not associated with any microscopic pathology changes. These weight effects were considered nonadverse and due to induction of liver metabolic enzymes. One male dog in the 40,000-ppm group demonstrated clinical signs, clinical pathology, and anatomic pathology changes consistent with canine juvenile polyarteritis syndrome (Clemo et al., 2003; Snyder et al., 1995); these effects were not considered to be test substance-related.

102.5.3  Chronic Toxicity and Oncogenicity The chronic toxicity and/or carcinogenicity Chlorantraniliprole were evaluated in rats and mice.

of

102.5.3.1  Rat, 2-Year Feeding In a 2-year combined chronic toxicity and oncogen­icity feeding study, Chlorantraniliprole was administered to male and female Crl:CD(SD)IGS BR rats (70 animals/sex/ concentration). A 1-year interim evaluation (10 animals/ sex/concentration) for chronic toxicity was included in the study. Concentrations in male and female rats were 0, 200, 1000, 4000, or 20,000 ppm (equivalent to 0, 7.7/11, 39/51, 156/212, or 805/1076 mg/kg bw/day males/females). No test substance-related effects on body weight, body weight gain, food consumption, or food efficiency were observed in male or female rats at any dietary concentration. Dietary expos­ ure to Chlorantraniliprole did not affect survival in male or female rats. No clinical or ophthalmological observations were attributed to treatment. There were no adverse changes in clinical pathology parameters, including hematology, coagulation, clinical chemistry, urinalysis, and urine corticosterone evaluations, in rats fed the test substance at dietary concentrations up to 20,000 ppm for 1 year, nor in WBC differential counts in rats fed the test substance at dietary concentrations up to 20,000 ppm for 2 years. An increase in liver weights was observed in female rats at 4000 ppm and 20,000 ppm (only at 1 year), but was not associated with any

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findings indicative of liver toxicity. Therefore, these weight changes were considered nonadverse and consistent with a pharmacological response to increased metabolism. No gross pathology findings were attributed to test substance exposure. Increased adrenal cortical microvesiculation due to lipid was present in the zona fasciculata region of the adrenal gland of some male rats in all dose groups in both the 1-year and 2-year studies. This finding was considered test substance-related but was not considered adverse as the adrenal morphology was generally in the range of what was observed in control rats, and the finding was not associated with any indication of cytotoxicity or other evidence of structural or functional impairment of the adrenal gland. No other treatment-related microscopic changes were observed in males or females. Chlorantraniliprole was not oncogenic in male or female rats.

102.5.3.2  Mouse, 18-Month Feeding In an 18-month carcinogenicity feeding study, Chlorantra­ niliprole was administered to male and female Crl:CD1(ICR) mice (70 animals/sex/concentration) at 0, 20, 70, 200, 1200, or 7000 ppm (equivalent to 0, 2.6/3.3, 9.2/12, 26/33, 158/196, or 935/1155 mg/kg bw/day males/females). Increased liver weights were observed in males and females at 1200 ppm, and increased hepatocellular hypertrophy was observed in males at 1200 ppm. Both the liver weight increases and hepatocellular hypertrophy were consistent with enzyme induction and were interpreted to not be adverse. Chlorantraniliprole was not oncogenic in male or female mice. Eosinophilic foci of cellular alteration were observed in the livers of a few 7000-ppm male mice at the termination of the study and were considered to be related to treatment. This change was interpreted in the study to be nonadverse due to its low incidence, the lack of an increase in liver tumors, the absence of degenerative changes in the liver in the affected mice, and the lack of similar findings in female mice. The study concluded the no-observedadverse-effect level (NOAEL) to be 7000 ppm (equivalent to 935 mg/kg bw/day in males and 1155 mg/kg bw/day in females) due to the absence of any adverse treatmentrelated effects. The participating regulatory agencies in the OECD global joint review of Chlorantraniliprole agreed the NOAEL in male mice should be established at 1200 ppm, equal to 158 mg/kg bw/day, based on the presence of eosinophilic foci, considered a potentially pren­eoplastic change, accompanied by hepatocellular hypertrophy and increased liver weight at 7000 ppm.

102.5.4  Effects on Adrenal Function In several studies conducted with Chlorantraniliprole in rats there was a slight increase in the degree of microvesiculation caused by lipid in the adrenal cortex of some animals. This histological observation in rats (graded minimal to

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mild) occurred almost exclusively in males following dietary administration of 90 days or longer and in a 28-day dermal toxicity study. No adrenal changes were observed in dogs or mice administered Chlorantraniliprole at high-dose levels for up to 12 and 18 months, respectively. The ad­renal cortex normally has a microvesicular appearance by light microscopy, which represents stores of lipid precursors for steroid hormone synthesis. The degree of adrenal cortical cell microvesiculation is variable and is influenced by a number of physiological processes that impact lipid storage (Guyton and Hall, 2002; Nussdorfer, 1986). The degree of microvesiculation seen in Chlorantraniliproletreated rats was within the range that may be observed histologically in control animals. Low incidences of minimally or mildly increased microvesiculation occurred in control rats in studies conducted with Chlorantraniliprole. To determine the toxicological significance of this finding in Chlorantraniliprole-treated rats, investigations were conducted to assess the structural and functional basis of the slight increases in lipid in adrenal cortical cells. Tissues were examined by electron microscopy from the adrenal cortex of selected male rats fed 0 or 20,000 ppm Chlorantraniliprole in the multigeneration reproduction study and at 1 year in the 2-year rat study. Histological grades of microvesiculation ranged from absent to mild. All cellular structures evaluated in the adrenal cortex were unaffected by treatment, no abnormal structures were observed in controls or treated rats, and the presence of lipid droplets was consistent with the normal morphology of the adrenal cortex. The functional impact of the increased degree of microvesiculation was evaluated by measuring corticoster­ one, the primary hormone produced by the adrenal cortex in rats, under both nonstressed (i.e., basal) conditions and under conditions of simulated physiological stress (i.e., adrenal cortical stimulating hormone- [ACTH]-induced). Basal corticosterone synthesis in rats administered Chlorantraniliprole was determined by measuring total corticosterone excreted overnight in urine 1 week prior to the sacrifice of all male and female rats designated for the 1-year interim sacrifice in the 2-year rat study. These rats had been fed Chlorantraniliprole up to a maximum dietary concentration of 20,000 ppm. Overnight urine corticoster­ one levels were similar across all dose groups including controls. Histologically, minimal to mild increases in the degree of lipid of the adrenal cortex was observed in some male rats in each of the Chlorantraniliprole-treated groups. The results indicated that Chlorantraniliprole does not affect basal corticosterone production in rats in the presence of slight increases in adrenal cortical cell microvesiculation. During conditions of physiological stress, the pituitary gland releases ACTH, which subsequently promotes the synthesis of glucocorticoids by the adrenal cortex. The ACTH stimulation test is a well-established clinical procedure in human and veterinary medicine that evaluates

Hayes’ Handbook of Pesticide Toxicology

the responsiveness of the adrenal cortex to ACTH. For this test, exogenous ACTH is administered to simulate conditions of physiological stress, and the adrenal cortical response is evaluated by measuring serum glucocorticoid concentrations. Corticosterone production in ACTH-stimulated rats was evaluated following dermal treatment with 1000 mg/kg bw/day Chlorantraniliprole for 28 days. In comparison with the control animals within the study, increased microvesiculation was present in several rats treated with Chlorantraniliprole. However, serum corticosterone concentrations were unaffected, indicating that Chlorantraniliprole did not inhibit corticosterone synthesis production under simulated stress conditions. In addition, lipid stores were proportionally decreased in control and treated animals following ACTH injection in com­ parison with control and treated rats in other studies without ACTH administration, indicating that lipid stores were appropriately mobilized for the synthesis of glucocorticoid hormones. This is considered further evidence that Chlorantraniliprole does not inhibit corticosterone synthesis in the adrenal cortex of rats. Thus, the slight increase in microvesiculation caused by lipid, which was observed microscopically in some male rats following exposure to Chlorantraniliprole, is a morphologic variation of what is observed in control animals. The histological change observed with Chlorantraniliprole administration was not associated with functional changes in the adrenal cortex or its capacity to respond to stress. Therefore, this change is considered to have no toxicological impact in studies conducted with Chlorantraniliprole.

102.5.5  Effects on Reproduction and Development The effects of Chlorantraniliprole on reproduction and development were investigated in developmental studies in both rat and rabbit and in a rat, two-generation reproduction study.

102.5.5.1  Rat, Developmental Toxicity In a developmental toxicity study, Chlorantraniliprole was administered by oral gavage to time-mated Crl: CD(SD)IGS BR female rats (22 animals/dose group) on gestation days 6–20. Gavage doses in 0.5% aqueous methylcellulose were 0, 20, 100, 300, or 1000 mg/kg bw/day. No test substance-related effects on maternal clinical observations, body weight, weight gain, food consumption, or gross post mortem observations were detected at any dose. Unscheduled maternal mortality did not occur. The mean number of corpora lutea, implantation sites, resorptions, live fetuses, fetal weight, and sex ratio were comparable across all groups. There were no test substance-related

Chapter | 102  Chlorantraniliprole: An Insecticide of the Anthranilic Diamide Class

fetal external, visceral, or skeletal malformations or vari­ ations observed at any dose.

102.5.5.2  Rabbit, Developmental Toxicity In a developmental toxicity study, Chlorantraniliprole was administered by oral gavage to time-mated Hra:(NZW)SPF female rabbits (22 animals/dose group) on gestation days 7–28. Gavage doses in 0.5% aqueous methylcellulose were 0, 20, 100, 300, or 1000 mg/kg bw/day. No maternal mortality occurred during the study. No test substance-related clinical observations were observed at any level. Maternal body weights, weight gain, and food consumption were similar across groups during the study. There were no gross postmortem maternal observations at any level. The mean number of corpora lutea, implantation sites, resorptions, live fetuses, fetal weight, and sex ratio were comparable across all groups. There were no test substance-related fetal external, visceral, or skeletal malformations or vari­ ations observed at any dose.

102.5.5.3  Rat, Two-Generation Reproduction In a two-generation reproduction study, Chlorantraniliprole was administered in the diet to male and female Crl: CD(SD)IGS BR rats (30 animals/sex/concentration for both the P1 and F1 generations). Concentrations were 0, 200, 1000, 4000, or 20,000 ppm (equivalent to 0, 12/16, 60/78, 238/318, or 1199/1594 mg/kg bw/day during premating in P1 males/females and 0, 18/20, 89/104, 370/406, or 1926/2178 mg/kg bw/day during premating in F1 males/ females). The P1 rats were bred within their treatment groups to produce F1 litters after at least 70 days on test. The F1 rats were bred within their respective treatment groups to produce F2 litters at least 82 days after weaning. There were no adverse test substance-related effects on body weight, weight gain, food consumption, or food efficiency, clinical signs of toxicity, or mortality in P1 and F1 males during premating and in P1 and F1 females during prema­ ting, gestation, or lactation. There were no test substancerelated effects on sperm motility, sperm morphology, and epididymal sperm or testicular spermatid numbers in either the P1 or F1 males at any dietary concentration. Similarly there were no effects produced by Chlorantraniliprole on the mean percent days in estrus, diestrus or proestrus, mean cycle length, or mean precoital interval in either the P1 or F1 females. Mating, fertility, gestation length, number of implantation sites, and implantation efficiency in either P1 or F1 generation were unaffected at any dietary concentration. An increase in mean liver weights was observed in P1 and F1 males and females at 4000 ppm and above, and was attributed to enzyme induction. In addition, a slight yet statistically significant increase in mean adrenal weight (absolute and/or relative to body weight) was observed at 4000- and 20,000-ppm P1 and F1 males and females. There were no adverse test substance-related effects on any

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gross or microscopic pathology endpoint. A minimal to mild increase in microvesiculation in the zona fasciculata region of the adrenal cortex was observed in some P1 adult males at 1000 ppm and greater, F1 adult males at 200 ppm and greater, and F1 adult females at 20,000 ppm. This finding was considered test substance-related but not adverse as the adrenal morphology was within the range that may be observed in control rats and was not associated with any indication of cytotoxicity or other evidence of structural or functional impairment of the adrenal gland. Mean body weight of the 20,000-ppm F1 pups was slightly reduced when compared with controls on lactation days 7, 14, and 21. The slightly lower 20,000-ppm pup weights were considered nonadverse as they were transient, small in magnitude, and F1 offspring weights were similar to controls by day 35 post-weaning. In addition, there were no effects on F2 offspring weights during lactation. A slight delay in the age of preputial separation and vaginal opening was observed in the 20,000-ppm F1 offspring. However, statistical analysis using body weight at time of weaning as a covariant demonstrated the lack of statistical significance. The delays were considered secondary to the slightly decreased body weight at the time of weaning and non­ adverse since the effect on body weight was transient.

102.5.6  Neurotoxic Effects The neurotoxic potential of Chlorantraniliprole was evaluated in rats in an acute oral dosing study and a 90-day feeding study.

102.5.6.1  Rat, Acute Oral Neurotoxicity In an acute neurotoxicity study, Chlorantraniliprole was administered to male and female Crl:CD(SD)IGS BR rats (12 animals/sex/dose) by single-dose oral gavage in 0.5% methylcellulose. Doses in male and female rats were 0, 200, 700, or 2000 mg/kg bw. There were no test substancerelated changes in body weight, weight gain, food consumption, food efficiency, mortality, clinical observations, forelimb or hindlimb grip strength, hindlimb foot splay, body temperature, rearing, duration of movement, number of movements, or any of the other behavioral parameters evaluated in a functional observational battery (FOB) in either males or females administered any dose of the test substance. In addition, there were no gross or microscopic test substance-related morphological changes in the ­nervous system tissues of these rats.

102.5.6.2  Rat, 90-Day Feeding Neurotoxicity In a 90-day neurotoxicity feeding study, Chlorantraniliprole was administered to male and female Crl:CD(SD)IGS BR rats (12 animals/sex/concentration). Concentrations administered to male and female rats were 0, 200, 1000,

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4000, or 20000 ppm (equivalent to 0, 13/15, 64/77, 255/304, or 1313/1586 mg/kg bw/day males/females). There were no test substance-related changes in body weight, weight gain, food consumption, food efficiency, mortality, clinical observations, forelimb or hindlimb grip strength, hind limb foot splay, body temperature, rearing, motor activity (duration of movement and number of movements), or any of the other behavioral parameters evaluated in the FOB in males or females at any dietary concentration. There were no test substance-related gross or microscopic morphological changes in the central or peripheral nervous system tissue or muscle tissue observed in males or females administered 20,000 ppm of the test substance.

102.5.7  Immunotoxic Effects In two 28-day immunotoxicity studies, Chlorantraniliprole was administered to male and female Crl:CD(SD)IGS BR rats and to Crl:CD1(ICR) mice (10 animals/sex/species/ concentration). Dietary concentrations in rats were 0, 1000, 5000, or 20,000 ppm (equivalent to 0, 74/82, 363/397, or 1494/1601 mg/kg bw/day males/females) and in mice were 0, 300, 1700, or 7000 ppm (equivalent to 0, 48/64, 264/362, or 1144/1566 mg/kg bw/day males/females). Prior to sacrifice, the immune system was stimulated by injecting sheep red blood cells (SRBC) on test day 22 (rats) or 23 (mice), and blood samples were collected from each animal on test day 28. The serum samples were assayed for their concentration of SRBC-specific IgM antibodies to provide a quantitative assessment of humoral immune response. Serum from animals similarly challenged with a positive control immunosuppressive agent were analyzed concurrently to provide confirmation that the assay performance was acceptable for detection of immunosuppression. There was no evidence of test substance-related toxicity

or immunosuppression in male and female rats or mice at any dietary concentration tested based on an evaluation of the in-life (body weight and food consumption parameters; clinical observations) and post-life (gross necropsy observations; absolute and relative brain, spleen, and thymus weights; and measurements of humoral immune response) data collected.

102.6  genotoxicity A battery of in vitro and in vivo tests was conducted to determine the genotoxic potential of Chlorantraniliprole. The tests included bacterial and mammalian gene mutation assays, the in vitro chromosome aberration assay, and the in vivo mouse micronucleus test. Negative results were obtained in all studies, indicating that Chlorantraniliprole does not cause genetic damage and therefore does not pose a mutagenic hazard (Table 102.2).

102.7  Toxicity to humans 102.7.1  Direct Observations and Health Records Chlorantraniliprole has been manufactured and marketed for a limited time. No reports of adverse health effects in manufacturing personnel or agricultural workers have been received.

102.7.2  Diagnosis of Poisoning No specific human symptoms of Chlorantraniliprole toxicity are known. Based on the results of animal tests, Chlorantraniliprole is not likely to be hazardous by the oral, dermal, or inhalation routes. No specific adverse

Table 102.2  Summary of Genotoxicity Studies with Chlorantraniliprole Type of study

Test system

Concentration/dose range tested

Result

In vitro bacterial mutagenicity (Ames)

Salmonella typhimurium and Escherichia coli

0, 2.5–5000 g/plate (with and without S-9)

Negative

In vitro chromosome aberration (clastogenicity)

Human lymphocytes

0, 125–500 g/mla (with and without S-9)

Negative

In vitro mammalian cell mutagenicity (CHO/HGPRT)

Chinese hamster ovary cells

0, 15.6–250 g/mlb (with and without S-9)

Negative

In vivo micronucleus

Mouse bone marrow

Male and female: 0, 500, 1000, 2000 mg/kg bw

Negative

a

Precipitate observed at 500 g/ml and above Precipitate observed at 250 g/ml and above

b

Chapter | 102  Chlorantraniliprole: An Insecticide of the Anthranilic Diamide Class

symptoms were noted in short- and long-term studies in laboratory animals. Given its low toxicity, no serious illness or mortality is expected from either accidental or prolonged overexposure to Chlorantraniliprole.

102.7.3  Sensitization Observations No cases of skin sensitization have been recorded in humans in the limited time that this product has been produced and marketed.

102.7.4  Proposed Treatment No specific antidotes or therapeutic regimens are indicated. No specific human effects of Chlorantraniliprole are known. Based on the results of animal tests, no specific intervention is indicated. If necessary, therapeutic efforts should be directed toward alleviation of any symptoms of illness.

102.8  Reference values and conclusions The OECD global joint review of Chlorantraniliprole in 2007 and the corresponding JMPR meeting in 2008 both concluded that this new active ingredient has no significant acute toxicity via the oral, dermal, and inhalation routes of exposure, is not a skin or eye irritant, and does not cause skin sensitization. It was agreed that Chlorantraniliprole is not carcinogenic, genotoxic, neurotoxic, or immunotoxic and does not cause reproductive or developmental toxi­ city. In short-term studies, the most consistent finding was a nonadverse induction of liver enzymes attributed to a pharmacological response to increased metabolism and a subsequent increase in liver weights. In male rats, a mild microvesiculation in cells of the adrenal cortex was within the range of that observed in control rats and did not impact cell structure or function in the numerous toxicology studies conducted and in specially designed adrenal function studies evaluating corticostersone concentrations in serum and urine. Therefore, the mild microvesiculation was concluded to have no toxicological significance. The liver effects observed in the 18-month mouse carcinogenicity study were agreed to form the basis for establishing the lowest NOAEL of 158 mg/kg bw/day due to the presence of eosinophilic foci of cellular alteration in a few male mice accompanied by hepatocellular hypertrophy and increased liver weight at the highest dose tested (935 mg/kg bw/day). On the basis of the available toxicology information, the following conclusions regarding reference values were made: Acute Reference Dose (ARfD): The establishment of an ARfD was not required for Chlorantraniliprole due to its low acute oral toxicity, the lack of developmental

l

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toxicity, and the absence of any other toxicological effects likely to be elicited by a single oral dose. l Acceptable Daily Intake (ADI) or Chronic Reference Dose (CRfD): An ADI or CRfD was established on the basis of eosinophilic foci accompanied by hepatocellular hypertrophy and increased liver weight in male mice in the 18-month feeding study for which a NOAEL of 158 mg/kg bw/day was concluded. Applying a safety factor of 100 to account for interspecies extrapolation and intraspecies variability, an ADI/CRfD of 1.58 mg/kg bw/ day was derived. The JMPR meeting noted that despite the absence of information on the chemical-specific mechanism of action with which to evaluate the relevance of the liver foci to humans, the observed foci may be a species- and sex-specific response and, therefore, of questionable toxicological significance. Therefore, the establishment of an ADI/CRfD on the basis of this endpoint was likely to have been conservative. l Occupational Endpoints: Based on the toxicological profile of Chlorantraniliprole, the U.S. Environmental Protection Agency and the Health Canada Pesticide Management Regulation Agency concluded that there were no identified hazards attributed to short- or intermediate-term exposure that required the establishment of a quantitative risk assessment for dermal and inhal­ ation exposures. Therefore, neither agency has established dermal or inhalation toxicity endpoints (U.S. EPA, 2008; Health Canada PMRA, 2008). In Europe, Chlorantraniliprole is presently being considered for approval by the European Commission. An acceptable occupational exposure level (AOEL) has been proposed by the lead reviewing authority (Ireland) based on the NOAEL (158 mg/kg bw/day) established in the 18-month mouse study. The AOEL is expressed as a systemic or internal dose; therefore the NOAEL was adjusted for the degree of oral bioavailability (13%) at a relevant dose (200 mg/kg bw) used in the rat metabolism study. With the application of a 100-fold safety factor for interspecies extrapolation and intraspecies variability, the proposed systemic AOEL is 0.2 mg/kg bw/day. For dermal expos­ ures, systemic doses were calculated by adjusting expos­ ures for absorption through the skin based on the degree of dermal penetration observed for Chlorantraniliprole in rats. A default assumption of 100% retention and absorption of inhaled product was assumed for inhalation expos­ ures. No risks were identified for pesticide applicators or for persons either adjacent to or re-entering treated areas. (Details contained in the Draft Assessment Report are not yet publically available.) The Australian Pesticides and Veterinary Medicines Authority concluded that the most appropriate study upon which to evaluate the risk to workers was the 28-day dermal toxicity study in rats in which the NOAEL was 1000 mg/kg bw/day, the highest dose tested. A margin of

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exposure approach was used in the calculation of risk to workers exposed to Chlorantraniliprole, and it was determined that personal protective equipment was not required (APVMA, 2008).

Acknowledgments The authors wish to thank Venkat Gaddamidi, Shawn Gannon, and Matthew Himmelstein for their scientific leadership in the conduct of the metabolism studies and metabolite identification in rats as well as the many other scientists who contributed to the toxicology studies conducted with Chlorantraniliprole.

References Australian Pesticides and Veterinary Medicines Authority. (2008). Public Release Summary on Evaluation of the New Active Chlorantraniliprole in the Products DuPont Coragen Insecticide, DuPont Altacor Insecticide, DuPont Acelepryn Insecticide. June 2008. http://www. apvma.gov.au/publications/downloads/prs_chlorantraniliprole.pdf. Clemo, F. A. S., Evering, W. E., Synder, P. W., and Albassam, M. A. (2003). Differentiating spontaneous from drug-induced vascular injury in the dog. Toxicol. Pathol. 31, 25–31. Cordova, D., Benner, E. A., Sacher, M. D., Rauh, J. J., Sopa, J. S., Lahm, G. P., Selby, T. P., Stevenson, T. M., Flexner, L., Gutteridge, S., Rhoades, D. L., Wu, L., Smith, R. M., and Tao, Y. (2006). Anthranilic diamides: A new class of insecticides with a novel mode of action, ryanodine receptor activation. Pest. Biochem. Physiol. 84, 196–214. Gutteridge, S., Caspar, T., Cordova, D., Rauh., J. J., Tao, Y., Wu, L., and Smith, R. M., World Patent Application WO2004027042. (2004). [US Patent Nos. 7,205,147 and 7,498,408 have been granted. Patents pending in other countries.] Guyton, A. C., and Hall, J. E. (2002). “Chapter 77: The Adrenocortical Hormones in Textbook of Medical Physiology,” 10th ed. Elsevier Press, Philadelphia, PA.

Hayes’ Handbook of Pesticide Toxicology

Health Canada Pest Management Regulatory Agency. (2008). Evaluation Report: Chlorantraniliprole. May 23, 2008. http://www.hc-sc.gc.ca/ cps-spc/pubs/pest/_decisions/erc2008-03/index-eng.php. Joint FAO/WHO Meeting on Pesticide Residues. (2008). Pesticide residues in food 2008, No. 193. Chlorantraniliprole, page 127–130. http://www.fao.org/fileadmin/templates/agphome/documents/Pests_ Pesticides/JMPR/JMPRReport08.pdf. Lahm, G. P., Selby, T. P., Freudenberger, J. H., Stevenson, T. M., Myers, B. J., Seburyamo, G., Smith, B. K., Flexner, L., Clark, C. E., and Cordova, D. (2005). Insecticidal anthranilic diamides: A new class of potent ryanodine receptor activators. Bioorg. Med. Chem. Lett. 15, 4898–4906. Lahm, G. P., Stevenson, T. M., Selby, T. P., Freudenberger, J. H., Cordova, D., Flexner, L., Bellin, C. A., Dubas, C. M., Smith, B. K., Hughes, K. A., Hollingshaus, J. G., Clark, C. E., and Benner, E. A. (2007). Rynaxypyr: A new insecticidal anthranilic diamide that acts as a potent and selective ryanodine receptor activator. Bioorg. Med. Chem. Lett. 17, 6274–6279. Nussdorfer, G. G. (1986). Chapter 6, Part 1A: Morphological-functional correlations in adrenocortical cells; The mechanism and the functional significance of the ultrastructural changes in adrenocortical cells; Lipid droplets in Volume 98: Cytophysiology of the Adrenal Cortex in International Review of Cytology, (G.H. Bourne and J.F. Danielli, eds.), Academic Press, Orlando, FL. Sattelle, D. B., Cordova, D., and Cheek, T. R. (2008). Insect ryanodine receptors: Molecular targets for novel pest control chemicals. Invert. Neurosi. 8, 107–119. Snyder, P. W., Kazacos, E. A., Scott-Moncrieff, J. C., HogenEsch, H., Carlton, W. W., Glickman, L. T., and Felsburg, P. J. (1995). Pathologic features of naturally occurring juvenile polyarteritis in beagle dogs. Vet. Pathol. 32, 337–345. U.S. Environmental Protection Agency. (2008). Chlorantraniliprole (DPX-E2Y45): Human health risk assessment for proposed uses on pome fruit, stone fruit, leafy vegetables, brassica leafy vegetables, cucurbit vegetables, fruiting vegetables, cotton, grapes, potatoes, turf and ornamentals. PC Code: 090100, Petition #7F7181 DP Barcode: D336983, D338120, D348103, D34624, Section 18 Registration #08LA01 (Rice), D346324. March 7, 2008. http://www.regulations. gov/fdmspublic/component/main?mainDocumentDetail&dEPAHQ-OPP-2007-0275-0004.

Section XV

Fumigants

(c) 2011 Elsevier Inc. All Rights Reserved.

Chapter 103

Sulfuryl Fluoride David L. Eisenbrandt1 and Jon A. Hotchkiss2 1 2

Dow AgroSciences, LLC, Indianapolis, Indiana The Dow Chemical Company, Midland, Michigan

103.1  Chemistry and formulations Sulfuryl fluoride, SO2F2, is manufactured and sold by Dow AgroSciences, LLC under the trade names Vikane gas fumigant and ProFume gas fumigant. The CAS registry number is 2699-79-8 and the molecular weight of sulfuryl fluoride is 102.07. Sulfuryl fluoride is an odorless, colorless gas with a melting point of 135.82°C and a boiling point of 55.38°C. The vapor pressure is 1.3  104 torr at 25°C. The solubility in water is 0.75 g/kg at 25°C. Sulfuryl fluoride is of low solubility in most organic solvents but is miscible with methyl bromide. Sulfuryl fluoride is not hydrolyzed by water, but is hydrolyzed by NaOH solution. Degradation of sulfuryl fluoride is via two-step hydrolysis (Figure 103.1) and requires a high pH. The half-life of sulfuryl fluoride in basic water (pH 9.2 at 20°C) is 3.3 min (Cady and Misra, 1974), while the half-life of fluorosulfate (FSO3) in basic water is 20 h (Jones and Lockhart, 1968). Sulfuryl fluoride is stable and noncorrosive by DOT definitions.

103.2  USES Since first marketed in 1961 as Vikane gas fumigant, sulfuryl fluoride has been used to fumigate over one million buildings, including houses, museums, historical landmarks, rare book libraries, government archives, and scientific and medical research laboratories (Dow AgroSciences, 1997). Sulfuryl fluoride is used to control a wide variety

Figure 103.1  Sulfuryl fluoride degradation is via two-step hydrolysis (base catalyzed). Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

of pests, including termites, cockroaches, clothes moths, bedbugs, rodents, and carpet beetles. Initial concentrations in fumigated structures are typically 2000–4000 ppm, although other concentrations may be used depending upon the target pest to be controlled, temperature, and the length of the exposure period. The activity of sulfuryl fluoride is dependent on the concentration reaching the target pest and the duration of exposure. Insect eggs require a higher dosage of sulfuryl fluoride compared to postembryonic life stages. Since the immature stages of some insects, such as termites and ants, cannot survive without adult care, dosages substantially less than required for egg control are effective. Sulfuryl fluoride is a restricted-use pesticide for use only by certified applicators. Because sulfuryl fluoride is a gas, structures are completely sealed prior to fumigation and a small amount of chloropicrin is introduced into the structure to warn people and animals that the structure is being fumigated (Dow AgroSciences, 1997). Chloropicrin has a noticeable disagreeable pungent odor at less than 1 ppm and causes irritation of the eyes, tears, and noticeable discomfort. ProFume gas fumigant is registered around the world for postharvest fumigation on stored commodities, such as cereal grains and dried fruit and tree nuts, as well as for pest control in food-handling establishments such as bakeries, food production facilities, mills, and warehouses. A wide spectrum of insect pests and all life stages of insects are controlled with this registered use of sulfuryl fluoride. Dow AgroSciences and ProFume gas fumigant have been recognized by the U.S. EPA as a partner in ozone protection and awarded the Stratospheric Ozone Protection Award. In addition, the company was awarded the Montreal Protocol Innovators Award for stratospheric ozone protection by the United Nations. The ACGIH Threshold Limit Value (TLV) and OSHA Permissible Exposure Level (PEL) for sulfuryl fluoride are 5 ppm TWA and 10 ppm STEL, respectively (ACGIH, 1998). 2245

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103.3  Hazard identification: toxicity to laboratory animals (Pre-1980) 103.3.1  Acute Toxicity Liquid sulfuryl fluoride directly contacting skin can result in frostbite (Torkelson et al., 1966). The 1-h LC50 in rats is between 2000 and 4000 ppm (Torkelson, 1959). Animals exposed to lethal doses of sulfuryl fluoride had tremors and convulsions. The convulsions were characterized by stiffening of the animal to a very rigid position and then toppling over backwards. Excessive salivation, loss of bladder control and chromodacryorrhea also were observed. A 1-h LC50 in male and female rats was 3020 and 3730 ppm, respectively (Vernot et al., 1977).

103.3.2  Subchronic Toxicity Lung, kidney, and liver microscopic changes were observed within 3 weeks in rats, guinea pigs, and rabbits exposed to 400 ppm sulfuryl fluoride for 7 h/day 5 days/week (Torkelson, 1959). In studies at 20, 50, 100, 150, and 200 ppm for up to 6 months, lung pathology was the most significant effect observed. In addition, slight central lobular degeneration and vacuolation of the liver and slight cloudy swelling of the tubular epithelium of the kidney were observed. Increased levels of fluoride in the blood, lungs, bone, and teeth as well as fluorosis of the teeth were observed. Groups of male and female rats (10/sex/dose) were maintained for 66 days on diets fumigated with 0 (control), 2, 10, 100, or 200 lbs/1000 ft3 of sulfuryl fluoride (Lockwood, 1958). The animals were weighed twice weekly for the first 28 days and then once a week thereafter and food consumption was recorded for the 1st month. They were observed frequently for gross changes in appearance or behavior and the teeth were examined for visual evidence of fluorosis. Samples of urine were obtained from all male rats for fluoride analysis. Terminal hematological values were obtained from five female rats in the 0-, 2-, and 10-lbs/1000 ft3 levels and from two male rats at each dietary level. Animals were autopsied at study termination and the lungs, heart, liver, kidneys, spleen, and testes were weighed. Portions of these organs as well as pancreas and adrenals were preserved and prepared for histological examination. Two additional rats of each sex were included in each dose group for fluoride analysis of blood, urine (male), kidney, lung, liver, and bones after 30 days on the test diets. Control diets averaged 36 ppm of fluoride, while diets fumigated with 2, 10, 100, or 200 lbs/1000 ft3 sulfuryl fluoride contained 55, 89, 386, or 740 ppm fluoride, respectively. Male and female rats tolerated a diet fumigated with sulfuryl fluoride at the rate of 2 lbs/1000 ft3 with no evidence of adverse effects, although fluoride content

Hayes’ Handbook of Pesticide Toxicology

of bone was increased somewhat. Diets fumigated with higher levels of sulfuryl fluoride resulted in retardation of growth and evidence of fluorosis in the teeth. The severity of the effects was directly proportional to the fluoride content of the diet. Fluoride analysis of urine and bones showed increased amounts of fluoride in proportion to the amount of sulfuryl fluoride exposure in the diet (in males, concentration of fluoride in bone was 260, 408, 413, 1615, and 1920 ppm and in urine was 9.9, 11.9, 13.3, 85.6, and 174.4 ppm for diets containing 36, 55, 89, 386, or 740 ppm fluoride, respectively); fluoride results for blood, kidney, lung, and liver were inconsistent with dosage levels.

103.4  Toxicity to laboratory animals (post-1980) 103.4.1  Acute Toxicity The 4-h LC50 in male and female Fischer 344 rats was 1122 and 991 ppm, respectively (Miller et al., 1980). Gross pathologic changes were observed primarily in the upper and lower respiratory tract. Histopathologic examination of animals exposed to 1200–1250 ppm revealed liver and kidney and possibly lung, heart, and spleen effects. Animals that survived for 2 weeks following exposure to 1200– 1250 ppm exhibited regenerative responses in the kidney. The highest concentration at which all animals survived was 450 ppm for males and 790 ppm for females. B6C3F1 mice were exposed to 400, 600 or 1000 ppm for 4 h (Nitschke and Lomax, 1989). All animals died within 90 min after termination of the exposure to 1000 ppm and within 6 days following exposure to 600 ppm. Body tremors were observed in several female mice shortly after exposure to 600 ppm and animals surviving after the exposure period were lethargic prior to death. There were no clinically visible effects noted in mice exposed to 400 ppm. In the B6C3F1 mouse, the 4-h LC50 was between 400 and 600 ppm for both males and females. CD-1 mice were exposed to 600, 700, or 800 ppm for 4 h to determine the LC50 (Nitschke and Quast, 1990). Effects noted in the CD-1 mouse were similar to those observed in the B6C3F1 mouse, but occurred at slightly higher concentrations. The 4-h LC50 was 660 and 642 ppm for males and females, respectively. Acute dermal exposure to sulfuryl fluoride was evaluated in rats exposed to 1000 or 9600 ppm for 4 h (Bradley et al., 1990). These concentrations are one and ten times greater than the 4-h LC50 in rats identified by Miller (1980). The exposures occurred in a Rochester-type chamber equipped with a door modified to allow the head of rats to protrude outside of the chamber. In the door, an elastic dental dam surrounded the neck of the rats and served as a barrier between the chamber air containing sulfuryl fluoride and the breathing air outside of the chamber. In addition, the backs of the rats were shaved prior to exposure to

Chapter | 103  Sulfuryl Fluoride

maximize dermal exposure. The only clinical effects during the exposure were chromodacryorrhea and fecal soiling. The incidence of chromodacryorrhea and fecal soiling was comparable between the two exposure groups and the effects were considered to be stress related, due to the method of restraint used during the exposure. There was no evidence of body tremors in these rats. Histopathologic examination of the skin and brain revealed no treatmentrelated lesions. Thus the dermal route does not appear to play a significant role in the toxicity of sulfuryl fluoride.

103.4.2  Acute Neurotoxicity Rats were exposed to 0, 100, or 300 ppm sulfuryl fluoride for 6 h/day for 2 consecutive days (Albee et al., 1993). A functional observational battery (FOB), grip performance, landing foot splay, motor activity, and a battery of electro­ diagnostic tests, including flash evoked potentials, somatosensory evoked potentials, and auditory brainstem responses, were conducted pre- and postexposure. Except for motor activity data, postexposure data were collected within 5 h after the second exposure. Motor activity data were collected 18 h after exposure. There were no exposure-related effects in any of the parameters.

103.4.3  Time to Incapacitation In an effort to understand the mode of action of sulfuryl fluoride, rats were exposed to 4000, 10,000, 20,000, or 40,000 ppm to determine the time to incapacitation (Nitschke et al., 1986). Rats were exposed to sulfuryl fluoride in a 14-l cylindrical chamber equipped with a motordriven activity wheel. Animals were forced to walk on the activity wheel for designated intervals during the exposure. Exposures were terminated when incapacitation or convulsions occurred. All rats either died or were moribund within 3 h following the end of the exposure. At the two highest concentrations, 20,000 and 40,000 ppm, rats were incapacitated within 12 min and died within 10 min after terminating exposure. At 10,000 ppm, rats were incapacitated after 16 min and, at 4000 ppm, rats were incapacitated after 40 min. At the lowest concentration, the mean survival time was 2.5 h after incapacitation occurred. Animals exposed to 10,000 ppm and higher appeared to be slightly cyanotic shortly after exposure occurred. The bluish skin discoloration disappeared within 10 min after purging the chamber with room air following exposure to 10,000 ppm. The skin discoloration did not appear to be reversible at higher concentrations. The cause of death at all concentrations appeared to be cardiovascular failure. Pulmonary congestion increased in severity as the concentration of the test chemical was increased in the atmosphere and at the two highest exposure concentrations, the pulmonary lesions appeared to contribute significantly to the death of the animals. At the two lowest concentrations, the pulmonary

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effects were not severe enough to be the major factor in the death of these animals. The lungs of rats exposed to 4000 or 20,000 ppm sulfuryl fluoride until incapacitation occurred were examined by light and electron microscopy (Eisenbrandt et al., 1987). Histopathologic examination of the lungs from rats exposed to 4000 ppm revealed minimal congestion, edema, and hemorrhage. Rats exposed to 20,000 ppm had more severe changes. Swelling and focal disruptions in alveolar epithelial cells were observed by electron microscopic examination of the lung. In addition, multifocal destruction of the alveolar wall with associated thrombosis was present in the lungs of rats exposed to 20,000 ppm. Both concentrations increased permeability of the alveolar wall as evidenced by the presence of edema, red blood cells, and fibrin in alveoli and interstitial spaces. The effects observed at the higher concentration, 20,000 ppm, appeared to play a significant role in the lethality of rats (Nitschke et al., 1986). The pulmonary effects noted at 4000 ppm were not severe enough alone to result in mortality. In another study seeking to understand the mode of action of sulfuryl fluoride, rats were exposed to 4000 or 10,000 ppm sulfuryl fluoride to determine the effect on respiration (Landry and Streeter, 1983). Respiratory frequency as well as tidal and minute volume were measured at 1-min intervals for 10 min prior to exposure and during a 20-min exposure. At 4000 ppm, there was a rapid initial increase in mean respiratory frequency and a decrease in mean tidal volume and mean minute volume relative to preexposure values. These respiratory parameters peaked after 2 min of exposure; there was a 39% increase in frequency, a 40% decrease in tidal volume, and a 23% decrease in minute volume. After approximately 10 min of exposure, frequency and tidal volume were near preexposure levels. These effects were considered to be insufficient to have resulted in mortality. Body temperature, heart rate, blood pressure, and electroencephalogram were monitored in rats exposed to 4000 or 20,000 ppm sulfuryl fluoride until rats expired (Gorzinski and Streeter, 1985). At 4000 and 20,000 ppm, animals survived for 79 and 14 min, respectively. A decrease in heart rate, a gradual increase in blood pressure, decreased respiration, occasional spiking and high-frequency loss in the EEG and power loss in the EEG were observed at both levels. All physiological parameters ceased to function at about the same time when the animals died, regardless of the concentration of sulfuryl fluoride, and did not help explain the cause of death in these animals.

103.4.4  Therapeutic/Amelioration of Toxicity Therapeutic treatment with calcium gluconate was evaluated because sulfuryl fluoride is extensively dehalogenated in termites (Meikle et al., 1963) and the similarity of effects in the mammalian incapacitation studies suggested fluoride

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toxicity (Nitschke et al., 1986). Rats were exposed to 4000 or 10,000 ppm sulfuryl fluoride for 45 or 16 min (sufficient to result in 100% lethality within 3 h), respectively, or treated i.p. with calcium gluconate prior to or after exposure to sulfuryl fluoride. Animals that were still alive 3 days after exposure to sulfuryl fluoride were considered to have survived. In separate groups of rats, serum fluoride levels were determined in rats exposed to sulfuryl fluoride alone or pretreated with calcium gluconate. Four of five rats pretreated with calcium gluconate prior to exposure to 4000 ppm sulfuryl fluoride survived 3 days after dosing. The other rat died 90 min after exposure. Treatment with calcium gluconate after exposure to sulfuryl fluoride was not effective. While survival was increased in rats pretreated with calcium gluconate, the surviving animals were extremely debilitated, with no apparent protection from convulsions. Animals pretreated with calcium gluconate did not survive exposure to 10,000 ppm sulfuryl fluoride. Administration of calcium gluconate resulted in an approximately 10% increase in serum calcium levels and did not appear to affect serum fluoride or magnesium levels (Nitschke et al., 1986). Rats were pretreated with one of three anticonvulsants, phenobarbital, diazepam, or diphenylhydantoin, prior to exposure to 4000 ppm sulfuryl fluoride or treated postexposure with phenobarbital and diazepam (Nitschke et al., 1986). These three anticonvulsants were selected due to their ready availability and different mechanisms of action. Phenobarbital administration before or after exposure to 4000 ppm sulfuryl fluoride for 45 min reduced the frequency and severity of convulsions and resulted in the survival of all animals. Treatment with phenobarbital following exposure of rats to 10,000 ppm sulfuryl fluoride for 15 min reduced the frequency of convulsions, but did not prevent death. Diazepam was less effective than pheno­ barbital and diphenylhydantoin did not have a beneficial effect. Thus, phenobarbital was most effective in ameliorating the acute toxic effects of overexposure to sulfuryl fluoride in rats.

103.4.5  Repeated Exposures The study design of several repeated exposure studies followed the appropriate EPA Test Guidelines (FIFRA Guideline Nos. 82-1, 82-4, 83-1, 83-2, and 835; U.S. Environmental Protection Agency, 1982). The study design for several of these studies is detailed by Eisenbrandt and Nitschke (1989) and Hanley et al. (1989). Briefly, animals were exposed to 99.8% pure sulfuryl fluoride diluted in filtered, humidified air for 6 h/day, 5 days/week for various time intervals. The analytical concentrations of sulfuryl fluoride in the air were measured by infrared spectrophotometry; analytical and nominal concentrations were in very close agreement. In general, animals were observed after exposure for changes in

appearance or behavior. Animals were weighed periodically. Blood samples were obtained for hematological and clinical chemistry determinations. Urine was collected from rats and various parameters were measured. Animals were sacrificed the day following the last exposure to sulfuryl fluoride. Terminal body weights and selected organ weights were recorded. All animals were examined for gross pathological alterations by a veterinary pathologist. Animals that died or were moribund prior to the scheduled sacrifice were necropsied as soon as possible. An extensive set of tissues was collected and processed for light microscopy by conventional techniques and stained with hematoxylin and eosin. In some cases special stains were used for selected tissues.

103.4.6  Subchronic Toxicity 103.4.6.1  Mice Groups of five mice/sex were exposed to 0, 30, 100, or 300 ppm sulfuryl fluoride for 6 h/day, 5 days/week for 2 weeks (Nitschke and Quast, 1995). All male mice and four of five female mice exposed to 300 ppm sulfuryl fluoride died during the 2nd week of the study. These animals lost weight and many had tremors. Vacuoles were observed in the cerebrum and/or medulla of eight of 10 mice exposed to 300 ppm and varied from very slight to moderate in severity. Four male and two female mice exposed to 100 ppm had very slight vacuoles in the cerebrum. The noobserved-effect level (NOEL) was 30 ppm. Groups of 14 mice/sex were exposed to 10, 30, or 100 ppm for 6 h/day, 5 days/week for 13 weeks (Nitschke and Quast, 1993). Four animals/sex/exposure level were used to measure serum fluoride levels. In addition, tissues of these four animals/sex/exposure level were perfused with glutaraldehyde/formaldehyde fixative and neural tissues were examined histopathologically. Standard parameters were evaluated in the main group of 10 mice/ sex/concentration. At the highest concentration, 100 ppm, there was approximately a 10% body weight decrease in male and female mice from control values. Except for elevated serum fluoride levels (Table 103.1), which followed a dose-response relationship, there were no exposure-related effects on clinical chemistry, hematology, urinalysis, organ weight, or gross pathology. Histopathologic examination of mice exposed to 100 ppm sulfuryl fluoride for 13 weeks revealed effects in the brain and thyroid gland (Tables 103.2 and 103.3). In the cerebrum, microvacuolation in male and female mice was observed in the external capsule and the caudate putamen. The effect was very slight to slight in severity and was bilaterally symmetrical. Microvacuoles also were observed in the region of the thalamus/hypothalamus of these animals. In this region, the microvacuoles usually extended from the external capsule and involved the adjacent amygdaloid region. There were no

Chapter | 103  Sulfuryl Fluoride

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recognizable inflammatory or degenerative changes associated with the microvacuoles. The single male mouse from the 100-ppm group without microvacuoles in the brain died during the course of the study due to accidental trauma. Histopathological changes in the brain of 4 species after 13 weeks of exposure to sulfuryl fluoride are compared in Tables 10.3.2 and 10.3.3. Microscopic changes in the thyroid gland were characterized by very slight hypertrophy of the

Table 103.1  Serum Fluoride Levels in Mice Following 13-Week Exposure to Sulfuryl Fluoride Concentration SO2F2, ppm

Males

Female

0

0.107  0.017

0.090  0.015

10

0.112  0.027

0.088  0.019

30

0.156  0.019

0.132  0.020a

100

0.259  0.073a

0.233  0.022a

N  4. a Statistically different from control mean by Dunnett’s test, alpha  0.05.

follicular epithelial cells associated with a decrease in the amount of colloid present. The NOEL was 30 ppm in mice exposed to sulfuryl fluoride for 6 hours/day, 5 days/week for 13 weeks.

103.4.6.2  Rats Groups of five animals of each sex were exposed to 100, 300, or 600 ppm sulfuryl fluoride for 6 h/day, 5 days/week for nine exposures in 2 weeks (Eisenbrandt and Nitschke, 1989). Nine of 10 rats in the 600-ppm exposure group were moribund and/or died between the second and sixth exposures. One female rat exposed to 600 ppm survived the nine exposures in the 2-week period. These animals had severe weight loss with body weights less than 70% of control values by the fifth exposure. Severe kidney lesions were observed in all rats exposed to 600 ppm. The papillary epithelium was necrotic over the tip of the papillae and the remainder of the epithelium was moderately hyperplastic. Subacute inflammation was associated with the necrosis and collecting ducts throughout the kidneys were dilated as a result of obstruction of the papillae. There was degeneration and necrosis of epithelial cells of the collecting ducts

Table 103.2  Histopathologic Observations in Brains of Males Exposed to Various Concentrations of Sulfuryl Fluoride for 13 Weeks Concentrationa

Species Mice

Control

Low

Middle

High

Brain (cerebrum), number of tissues examined

10

10

10

10

Vacuolation caudate putamen very slight

0

0

0

7

Vacuolation caudate putamen slight

0

0

0

2

Vacuolation external capsule very slight

0

0

0

7

Vacuolation external capsule slight

0

0

0

2

Brain (cerebrum), number of tissues examined

10

10

10

10

Vacuolation cerebrum slight

0

0

0

10

Brain (cerebrum), number of tissues examined

7

7

7

7

Malacia severe

0

0

0

3

Vacuolation very slight

0

0

0

3

Brain (midbrain), number of tissues examined

4

4

4

4

Vacuolation very slight

0

0

0

1

Rats

Rabbits

Dogs

a

Animals were exposed to sulfuryl fluoride for 6 h/day, 5 days/week for 13 weeks. The low, middle, and high concentrations corresponded to targeted concentrations of 10, 30, and 100 ppm in mice; 30, 100, and 300 ppm in rats; 30, 100, and 300 ppm in rabbits; and 30, 100, and 200 ppm in dogs, respectively.

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Table 103.3  Histopathologic Observations in Brains of Females Exposed to Various Concentrations of Sulfuryl Fluoride for 13 Weeks Concentrationa

Species Mice

Control

Low

Middle

High

Brain (cerebrum), number of tissues examined

10

10

10

10

Vacuolation caudate putamen very slight

0

0

0

3

Vacuolation caudate putamen slight

0

0

0

5

Vacuolation external capsule very slight

0

0

0

6

Vacuolation external capsule slight

0

0

0

4

Brain (cerebrum), number of tissues examined

10

10

10

10

Vacuolation cerebrum slight

0

0

0

10

Brain (cerebrum), number of tissues examined

7

7

7

7

Malacia severe

0

0

0

1

Vacuolation very slight

0

0

0

3

Vacuolation slight

0

0

0

2

Vacuolation moderate

0

0

1

0

Brain (midbrain), number of tissues examined

4

4

4

4

Vacuolation very slight

0

0

0

1

Rats

Rabbits

Dogs

a

Females were exposed to same concentrations as males in Table 103.2.

with regeneration of surviving epithelial cells. Several rats had mineralization at the junction of the inner and outer medulla, which was accompanied by chronic interstitial inflammation in some animals. Degenerative and regenerative changes in the proximal tubules were also observed. Minimal kidney changes were observed in rats exposed to 300 ppm; there were no exposure-related kidney lesions observed in rats exposed to 100 ppm. The sole surviving female rat exposed to 600 ppm sulfuryl fluoride had severe, diffuse inflammation and multifocal ulceration of the nasal mucosa as well as slight bronchioalveolar inflammation in the lungs. Several of the rats that died after exposure to 600 ppm sulfuryl fluoride had pulmonary edema and/ or hemorrhage or fibrin within alveoli. The NOEL was 100 ppm in rats exposed to sulfuryl fluoride for 2 weeks. Groups of 10 rats/sex were exposed to 0, 30, 100, or 300 ppm for 6 h/day, 5 days/week for 13 weeks (Eisenbrandt and Nitschke, 1989). Effects attributed to sulfuryl fluoride toxicity included mottled teeth in rats exposed to 100 or 300 ppm, decreased specific gravity of the urine, and histopathological changes in the respiratory tract, brain, and kidneys of rats exposed to 300 ppm. The

respiratory tract effects consisted of very slight to severe inflammation in the nasal tissue with mucopurulent exudate in the nasal passages in the more severe cases. The more extensive inflammation was accompanied by degeneration and reactive changes in the mucosa. Slight subpleural histiocytosis also was observed in the lungs of rats exposed to 300 ppm sulfuryl fluoride. In the brain, minimal vacuolation in the area of the caudate-putamen nuclei was observed in rats exposed to 300 ppm and was more prominent in the white fiber tracts of the internal capsule than in the adjacent neuropil (Tables 103.2 and 103.3). Special stains of the brain with LFBPAS or Sevier Munger stain did not reveal any additional effects. Very slight hyperplasia of the renal collecting ducts was most apparent in the outer portion of the inner zone of the medulla of most female rats exposed to 300 ppm. As a separate part of the above-mentioned 13-week study, groups of seven rats/sex were exposed to 0, 30, 100, or 300 ppm for the same time period for the specific purpose of evaluating neurological function (Mattsson et al., 1988). After 13 weeks of exposure to sulfuryl fluoride, hindlimb grip strength, observational battery, visual

Chapter | 103  Sulfuryl Fluoride

evoked response, cortical flicker fusion, auditory brainstem response to tone pips, auditory brainstem response to clicks, cerebellar evoked response, somatosensory evoked response, and caudal nerve action potential were measured approximately 12 h after the last exposure to sulfuryl fluoride. All but two males and two females from the 0- and 300-ppm exposure groups were subjected to a gross pathologic examination at the end of the 13-week study. The two animals/sex from the 0- and 300-ppm exposure group were evaluated with an auditory brainstem response approximately 8 weeks after the last exposure to sulfuryl fluoride and then sacrificed. Hindlimb grip strength was normal for all rats, but evoked responses were clearly altered at 300 ppm and slightly altered at 100 ppm. The principal effect was decrease in flicker fusion (CFF) and a slowing of all waveforms at 300 ppm, as well as a slowing of the visual evoked response and the somatosensory evoked response of female rats at 100 ppm. Histopathologic changes in the brain consisted of vacuoles in the white fiber tracts of the caudate putamen. No necrosis or neuronal destruction were noted. The two rats exposed to 300 ppm in the recovery group had normal auditory brainstem responses and normal brain histopathology. The fact that all of the evoked responses in rats exposed to 300 ppm for 13 weeks were affected suggested a widespread functional CNS effect. The electrophysiologic slowing of the evoked responses were felt to be due to some mechanism other than the minor vacuolization observed in the caudate putamen. The NOEL was 30 ppm in rats exposed to sulfuryl fluoride for 13 weeks in the standard subchronic and neurological studies.

103.4.6.3  Rabbits Groups of three rabbits of each sex were exposed to 100, 300, or 600 ppm sulfuryl fluoride for 6 h/day, 5 days/week for nine exposures in 2 weeks (Eisenbrandt and Nitschke, 1989). All rabbits exposed to 600 ppm were hyperactive and one animal had a convulsion that resulted in a fractured tibia. A second rabbit had a fractured vertebra that may have been the result of a convulsion. Treatment-related malacia (necrosis) was present in the cerebrum of all rabbits exposed to 600 ppm and one male and one female rabbit exposed to 300 ppm. Reactive gliosis and demyelination accompanied the malacia. Rabbits exposed to 300 or 600 ppm also had vacuoles in the globus pallidus and putamen as well as the external and internal capsules of the brain. Moderate, subacute to chronic inflammation of nasal tissues with mucopurulent exudate in the nasal cavities was observed in most rabbits exposed to 300 or 600 ppm sulfuryl fluoride. The inflammation was probably due to irritation of the nasal mucosa due to the test material. At the higher concentration, acute inflammation was observed in the trachea, bronchi, and bronchioles of some rabbits and

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may have been treatment-related. The NOEL was 100 ppm in rabbits exposed to sulfuryl fluoride for 2 weeks. In a 13-week study, groups of seven rabbits/sex were exposed initially to 30, 100, or 600 ppm for 6 h/day, 5 days/ week (Eisenbrandt and Nitschke, 1989). After 2 weeks exposure to 600 ppm, the target concentration was reduced to 300 ppm, which resulted in an average concentration over the 13-week period of 337 ppm. The exposure concentration was reduced from 600 ppm to 300 ppm due to convulsions observed in one male and one female. A second female rabbit exposed to 600 ppm was euthanized after eight exposures due to a fractured vertebra. No clinically visible effects were noted in rabbits exposed to 300 ppm or lower concentrations. Except for elevated serum fluoride levels that followed a dose-response relationship, there were no clinical chemistry, hematology, urinalysis, organ weight, or gross pathologic changes observed. Histopathologic changes were noted in the nasal tissues and brain of rabbits exposed to 300 and in the nasal tissue of one male and brain of one female exposed to 100 ppm (Tables 103.2 and 103.3). In the nasal tissues, varying degrees of purulent nasal exudate, olfactory epithelial degeneration and hyperplasia, and hypertrophy of the respiratory epithelium in the nasal turbinates were observed. The brain changes consisted of vacuolation of the white matter at 100 ppm. In rabbits exposed to 300 ppm, malacia of the internal and external capsules, putamen, and globus pallidus were observed. Some animals exposed to 300 ppm had gliosis and/or hypertrophy of vascular endothelial cells in the same area. Special stains of the brain with LFB-PAS or Sevier Munger stain were not remarkable. The NOEL was 30 ppm in rabbits exposed to sulfuryl fluoride for 13 weeks.

103.4.6.4  Dogs Groups of one male and one female dog were exposed to target concentrations of 0, 30, 100, or 300 ppm for 6 h/day 5 day/week for nine exposures (Nitschke and Quast, 1991). At 300 ppm, infrequent intermittent episodes of tremors and tetany were observed in both dogs beginning with the fifth exposure. On test day 9, during the seventh exposure, the tremors and tetany were severe enough that the exposure was terminated after approximately 5.5 h. Within 30 min after terminating the exposure, both dogs appeared to be normal. Similar clinical effects were noted during subsequent exposure periods and were rapidly reversible even during the exposure period. There were no exposurerelated clinical effects in dogs exposed to 30 or 100 ppm. The female dog exposed to 300 ppm sulfuryl fluoride lost approximately 500 g body weight over the nine exposures. Serum fluoride levels of dogs exposed to 100 or 300 ppm were approximately two- to fourfold higher than control values. Serum calcium levels measured shortly after exposure on test days 5 and 9 when the animals appeared to be clinically normal were comparable to control levels. There

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were no exposure-related hematological, organ weight, or gross pathological effects noted in dogs exposed to concentrations as high as 300 ppm. Minimal microscopic inflammatory changes were observed in the nasal turbinates of the male and female dog and trachea of the female dog exposed to 300 ppm. Although numerous microscopic sections were examined from the cerebral cortex, brainstem, cerebellum, and medulla oblongata, there were no histopathologic changes detected in dogs exposed to 300 ppm. The NOEL was 100 ppm in dogs exposed to sulfuryl fluoride for 2 weeks. In a 13-week study, groups of four male and four female beagle dogs were exposed to 0, 30, 100, or 200 ppm sulfuryl fluoride for 6 h/day, 5 days/week for 13 weeks (Nitschke et al., 1992). One male dog exposed to 200 ppm was laterally recumbent with tetany, tremors, salivation, and incoordination 75 min after exposure on test day 19. One hour later, the activity of this animal was decreased relative to controls but was otherwise normal. Similar effects were not observed during the remainder of the study. After 13 weeks of exposure to sulfuryl fluoride, the mean body weight value of male and female dogs exposed to 200 ppm were 88 and 96%, respectively, of control values. Mean body weight values of male and female dogs exposed to lower concentrations of sulfuryl fluoride were comparable to control values. There were no exposurerelated hematological, clinical chemistry, urinalysis, organ weight, or gross pathological effects. Histopathologically, a single small bilaterally symmetrical focal change was noted in the putamen of the midbrain of one male and one female dog exposed to 200 ppm (Tables 103.2 and 103.3). The minimal focal change was characterized by vacuolation, gliosis, perivascular cuffing, and hypertrophy of endothelial cells, and individual cells showed nuclear pyknosis and karyorrhexis. The focal reaction was slightly more prominent in the male compared to the female. All other microscopic observations were considered to be incidental findings unrelated to exposure to sulfuryl fluoride. The NOEL was 100 ppm in dogs exposed to sulfuryl fluoride for 6 h/day, 5 days/week for 13 weeks.

103.4.7  Chronic Toxicity 103.4.7.1  Mice Groups of 50 mice/sex were exposed to measured vapor concentrations of 0, 5, 20, or 80 ppm sulfuryl fluoride for 6 h/day, 5 days/week for 18 months (Quast et al., 1993b). Ten additional mice/sex/exposure level were randomly designated as a satellite group for necropsy after 12 months of exposure to evaluate chronic toxicity. During the 1st year of exposure, a slightly earlier onset of mortality was observed in the 80-ppm males; however, mortality at the end of the 18 months of exposure was not statistically identified as increased in any of the sulfuryl

Hayes’ Handbook of Pesticide Toxicology

fluoride-exposed groups of male mice. The female mortality rate during the 1st year of exposure was comparable in all groups. In both males and females exposed to 80 ppm, the incidence of exudative rhinitis and aspiration pneumonia accompanied by an impacted esophagus was increased from control values. Body weight of 80-ppm male mice was 10% higher than control values after 6 months and body weights of female mice exposed to 80 ppm were significantly decreased from control values after 1 month exposure. In general, the effect in the female mice was not as severe as in the male mice during the first 12 months. During the last several months of exposure, the body weight difference between control and high-dose female mice ranged from 10 to 15%. The body weights of male and female mice exposed to 5 or 20 ppm were comparable to control values throughout the study. Clinical chemistry and hematology values of male and female mice exposed to the various concentrations of sulfuryl fluoride were comparable to control values at the 12-month and terminal sacrifice. Although terminal body weight effects were observed in male and female mice exposed to 80 ppm sulfuryl fluoride for 12 or 18 months, the organ weight effects observed were considered to be due to the marked body weight differences and not indicative of target organ toxicity. There were no treatment-related gross pathological effects noted in mice at 12 months. There was a decreased incidence of normally occurring spontaneous gross lesions observed in animals exposed to 80 ppm and included a marked decrease in the incidence of cystic ovaries and cystic endometrial hyperplasia of the uterus in females and a decreased incidence of dilated kidney pelvis. Based upon the necropsy findings, there were no target organs identified in the mice after 12 or 18 months. At the 12-month interim sacrifice, histopathologic examination of male and female mice exposed to various concentrations of sulfuryl fluoride revealed changes in the brain and thyroid gland of animals exposed to 80 ppm only. Essentially all 80-ppm exposed mice had very slight or slight microscopic vacuolation of the cerebrum in the region of the external capsule. The caudate putamen was only affected in one male mouse in the 80-ppm group. The amygdaloid adjacent to the external capsule was not affected. The vacuolation in the cerebrum was suggestive of edematous change and was not associated with an inflammatory cell reaction. Very slight hypertrophy of the thyroid follicular epithelial cells was observed. More male mice were affected than female mice. Interestingly, mice exposed to 80 ppm for 12 or 18 months exhibited a lower incidence and severity of brain effects than mice exposed to 100 ppm for 13 weeks (Nitschke and Quast, 1993). In the oncogenicity group, target organs were limited to those previously defined after 12 months and consisted of the brain and thyroid gland of mice exposed to 80 ppm

Chapter | 103  Sulfuryl Fluoride

only. Only a quarter of the mice exposed to 80 ppm sulfuryl fluoride had histopathologic changes in the external capsule of the brain. There were no recognizable changes in the caudate putamen or amygdaloid regions of the brain. Thyroid changes in mice exposed to 80 ppm were characterized by hypertrophy of follicular epithelial cells. These thyroid changes occurred at a much lower incidence in mice at 18 months when compared to 12 months, with males having a higher incidence than females. All other microscopic changes were considered to be unrelated to sulfuryl fluoride exposure. There was no increase in the incidence of any tumor in male or female mice exposed to concentrations as high as 80 ppm sulfuryl fluoride.

103.4.7.2  Rats Groups of 50 male and 50 female rats were exposed to 0, 5, 20, or 80 ppm sulfuryl fluoride for 6 h/day, 5 days/week for 2 years (Quast et al., 1993c). Fifteen additional rats/sex/ exposure level were randomly designated at the beginning of the study as a satellite group for assessment of general toxi­ city and neurotoxicity (functional observational battery, motor activity, and perfusion-fixed histopathology of nervous system tissues) following 1 year of exposure. Mortality of rats exposed to sulfuryl fluoride through the first 16 months of the study was similar to the control group in both male and female rats. After 16 months of exposure to 80 ppm, mortality of both males and female rats was increased from control values. The mortality rate of female rats exposed to 5 or 20 ppm was lower than control values from 20 months until the end of the study. Slight body weight effects were observed in female rats exposed to 80 ppm throughout the 1st year; body weights of male and female rats exposed to 80 ppm became progressively more severe after 1 year. There were no consistent effects observed during the 1st year in urinary specific gravity values of males or females exposed to sulfuryl fluoride. However, urinary specific gravity values of male and female rats exposed to 80 ppm for 19 or 21 months were significantly decreased from control values. Several clinical chemistry parameters commonly associated with kidney toxicity were affected in rats exposed to 80 ppm for 19 and 21 months. These included increased urea nitrogen, cholesterol, triglycerides, creatinine, and phosphorus and decreased total protein, albumin, and chloride. In addition, albumin levels of male rats exposed to 80 ppm for 12 months were also affected. At the 12-month sacrifice, increases in the relative kidney and liver weights of male rats exposed to 80 ppm were the only organ weight differences noted in rats exposed to sulfuryl fluoride. Histopathologic changes were noted in the kidneys, lungs, and teeth of rats exposed to 80 ppm sulfuryl fluoride for 12 months. A very slight to slight degree of chronic progressive glomerulonephropathy was noted in males and females, with females generally less severely affected. In the lungs, very slight to slight aggregates of alveolar

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macrophages were noted in rats. These effects were very minimal and were not considered to significantly impair pulmonary function. Very slight to slight dental fluorosis involving the upper incisor teeth was also observed. The molars of these animals were unaffected. There were no exposure-related histopathologic changes noted in rats exposed to 5 or 20 ppm. As in rats exposed to sulfuryl fluoride for 12 months, gross and histopathologic changes were noted in the kidneys, lungs, and teeth of rats exposed to 80 ppm sulfuryl fluoride for a lifetime. While there was no apparent difference in the gross and histopathologic changes noted in the lungs and teeth of rats exposed for 12 or 24 months, the kidney changes had progressed from very slight or slight to severe or very severe chronic progressive glomerulonephropathy. These kidney changes have been commonly observed in control animals from lifetime studies; however, the incidence rate was much higher in the high-exposure animals. Along with the kidney changes, secondary changes, such as hyperparathyroidism and mineralization of many tissues were observed. These secondary changes were described previously by Boorman et al. (1990) and Mohr et al. (1992). Except for a very slight fluorosis of the teeth of male rats exposed to 20 ppm, there were no effects observed in male or female rats exposed to 5 or 20 ppm. There was no increase in the incidence of any tumor in male or female rats exposed to concentrations as high as 80 ppm sulfuryl fluoride. There was no evidence of a nervous system effect based on functional observational battery, motor activity, or histopathological examination of perfusion-fixed nervous system tissues (Spencer et al., 1994). The NOAEL for chronic toxicity was 20 ppm in males due to several rats with very slight fluorosis. The NOEL was 5 ppm in males and 20 ppm in females.

103.4.7.3  Dogs Groups of four dogs/sex were exposed to 0, 20, 80, or 200 ppm for 6 h/day, 5 days/week for 1 year (Quast et al., 1993a). Body weight gains in the highest exposure group of males and females were less than controls within the first 2 weeks of exposure to sulfuryl fluoride and the differences became greater throughout the study until the dogs were removed due to morbidity or death. Although no clinical effects were noted in dogs exposed to 200 ppm for the first 8 months, clinical effects were observed in these animals at approximately 9 months into the study. Observations in these animals included labored breathing, shallow, rapid respiration, and pale or blue mucous membranes. The onset of these observations was relatively swift, in the first dog, effects were noted on test day 263 and the animal died on test day 267. Due to the relatively swift onset, the last dog was sacrificed on test day 282 when the exposure was stopped due to excessive toxicity.

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There were no exposure-related effects noted in dogs exposed to 20 or 80 ppm sulfuryl fluoride for 1 year or dogs exposed to 200 ppm through 6 months. Effects were noted in hematological or clinical chemistry values of dogs exposed to 200 ppm for approximately 9 months. However, by this point, these dogs were starting to show signs of respiratory distress and the observed effects were considered to be minor secondary hematological and clinical chemistry changes. Gross examination of dogs exposed to 200 ppm revealed dark lungs, which appeared to be consolidated. There were no other tissues affected in dogs exposed to 200 ppm for approximately 9 months or dogs exposed to concentrations of 20 or 80 ppm for 1 year. Histopathologic changes were noted in the lungs, brain, thyroid gland, and canine teeth of dogs exposed to 200 ppm and in the lungs and canine teeth of dogs exposed to 80 ppm. The pulmonary changes appeared to be a chronic active inflammation, which primarily involved the peripheral regions of the lung of animals exposed to 200 ppm without recognizable alterations in the major airways. An increased number of alveolar macrophages was observed in scattered alveoli. In the more advanced stages of the chronic active inflammation process, these foci apparently increased in size and hypertrophied type II pneumo­ cytes were observed. In addition, epithelial cells were hypertrophied and hyperplastic. In the more severe cases, a focal thickening of the pleura and interalveolar septa was also observed. Dogs exposed to 80 ppm had a very slight increase in the aggregates of alveolar macrophages, with several dogs exhibiting a very slight degree of the chronic active inflammation. In the brain, a focus of malacia was observed in five of eight dogs inhaling 200 ppm. Although inflammatory cells were noted, they appeared to be an insignificant factor. Very slight hypertrophy of the follicular epithelium was observed in the thyroid gland of all male and three female dogs. Since the dogs were approximately 5 months of age when initially exposed to sulfuryl fluoride, the teeth of these animals were still growing. Consequently, during the beginning of the study, concentric rings were observed for each exposure period. Thus five concentric rings were observed for each week of exposure. However, as these animals matured it was more difficult to recognize the concentric rings. There were no exposure-related effects noted in dogs exposed to 20 ppm sulfuryl fluoride.

103.4.8  Teratology Studies 103.4.8.1  Rats Groups of 35–36 bred Fischer 344 rats were exposed to 0, 25, 75, or 225 ppm sulfuryl fluoride for 6 h/day on days 6–15 of gestation (Hanley et al., 1989). There was no evidence of embryotoxicity, fetotoxicity, or teratogenicity noted in

rats exposed to concentrations as high as 225 ppm sulfuryl fluoride.

103.4.8.2  Rabbits Groups of 28–29 inseminated New Zealand White rabbits were exposed to 0, 25, 75, or 225 ppm sulfuryl fluoride for 6 h/day on days 6–18 of gestation (Hanley et al., 1989). Pregnant rabbits exposed to 225 ppm lost weight during the exposure period and weight gain during the postexposure period (days 19–29 of gestation) was also decreased significantly. Body weights of rabbits exposed to 25 or 75 ppm were unaffected. Body weights of the fetuses from dams exposed to 225 ppm were significantly lower (14% decrease) than in the control group. Fetal crown-rump length also was decreased slightly in this group. There was no evidence of embryotoxicity, fetotoxicity, or teratogeni­ city noted in rabbits exposed to concentrations as high as 225 ppm sulfuryl fluoride.

103.4.9  Reproduction Toxicity 103.4.9.1  Rats Groups of 30 male and 30 female Sprague-Dawley rats were exposed to 0, 5, 20, or 150 ppm sulfuryl fluoride for 6 h/day, 5 days/week for 10 weeks for the F0 and 12 weeks for the F1 generation prior to mating and 6 h/day, 7 days/ week during mating, gestation, and lactation through two generations (Breslin et al., 1993). Body weights of F0 male and female rats exposed to 150 ppm were significantly decreased from control values during most of the premating period. The body weight gain of these female rats during gestation also was decreased. However, during the lactation period, the body weight gain was increased, possibly as a compensatory mechanism. There was no exposure-related effect on the F0 male or female conception index, fertility indices, length of gestation, time to mating, pup survival indices, or pup sex ratio. Similarly, there was no exposure-related effect on the number of F1 pups born dead or alive, or on the litter size at any exposure level. However, the litter size of live pups at birth and on days 1 and 4 before culling were increased 2.0 pups/litter in the 5-ppm exposure group and 1.5 pups/litter in the 20-ppm exposure group when compared to the control group. While these increases in litter size were not considered exposure-related, as a dose–response was not observed and no effects on litter size were observed in animals exposed to 150 ppm, the increase in litter size above control values has biological significance in that average pup weights are known to decrease with increasing litter size. Indeed, F1 female pup body weights were occasionally decreased in the 5-ppm exposure group. On the other hand, the body weights of F1 male and female pups from dams exposed to 150 ppm were statistically decreased throughout

Chapter | 103  Sulfuryl Fluoride

most of the lactation period. While the decreases in body weight of F1 pups from the 150-ppm exposure group were attributed to treatment, these body weight effects were considered secondary to the decreased maternal growth observed throughout the premating and gestation periods. Body weight effects in the F1 adults were similar to that observed in the F0 generation. Exposure-related effects were noted in body weights of male and female F1 rats exposed to 150 ppm during premating and female F1 rats exposed to 150 ppm during gestation and lactation but not at lower exposure levels. No exposure-related effects were observed on the F1 male or female fertility indices, length of gestation, time to mating, pup survival indices, or pup sex ratio in any exposure group. Similarly, no exposure-related effects on the number of F2 pups born alive or dead, or on the litter size were observed at any exposure level. No exposure-related effects on the body weights of male or female F2 pups from dams exposed to 5 or 20 ppm were observed at any time during the lactation period. However, body weights of male and female pups from dams exposed to 150 ppm were significantly decreased on lactation days 14 and 21. This was considered to be secondary to the decreased maternal growth observed during the premating and gestation periods. The decreased growth of F2 pups from dams exposed to 150 ppm sulfuryl fluoride was less severe than the decreased weights observed in the F1 pups at the same exposure level. The pathologic changes in the teeth, lungs, and brain in the adult F0 and F1 Sprague-Dawley rats in this study were essentially identical to those observed in a subchronic study in Fischer 344 rats previously mentioned (Eisenbrandt and Nitschke, 1989). Minor changes in the lung and brain were present at lower exposure concentrations in this study (20 ppm in the lung and 150 ppm in the brain) than previously noted in Fischer 344 rats. Interestingly, the effects observed in the brain of the F1 adults occurred in fewer animals than in the F0 adults even though the length of exposure to sulfuryl fluoride was increased. Female rats had a higher incidence of brain and lung lesions than did males at a given exposure level. The parental NOEL was 5 ppm, the NOEL for neonatal growth was 20 ppm, and the NOEL for reproductive toxicity and fertility was 150 ppm.

103.4.10  Genetic Toxicity 103.4.10.1  Ames Test Sulfuryl fluoride was tested in strains TA98, TA100, TA1535, TA1537 with and without metabolic activation (Gollapudi et al., 1990b). Petri plates were exposed for 4 h at 37°C to nominal concentrations of 300, 1000, 3000, 10,000, and 30,000 ppm sulfuryl fluoride. The plates were incubated for an additional 2 days prior to determining the frequencies of mutants/plate. Sulfuryl fluoride was not mutagenic in any of the tester strains.

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103.4.10.2  Unscheduled DNA Synthesis The genotoxicity of sulfuryl fluoride was evaluated in the rat hepatocyte unscheduled DNA synthesis (UDS) assay (Gollapudi et al., 1991). In two separate assays, sulfuryl fluoride did not elicit a positive UDS response at nominal concentrations ranging from 204 to 1020 ppm.

103.4.10.3  Micronucleus Test Sulfuryl fluoride was evaluated in the mouse bone marrow micronucleus test (Gollapudi et al., 1990a). Groups of mice were exposed to 0, 50, 175, or 520 ppm sulfuryl fluoride. Mice were sacrificed at 24, 48, or 72 h after exposure to sulfuryl fluoride. There were no significant increases in the frequencies of micronucleated polychromatic erythrocytes in the bone marrow of mice.

103.4.11  Uptake and Metabolism Studies in termites have demonstrated extensive dehalo­ genation of 35S-labeled sulfuryl fluoride (Meikle et al., 1963). The fluoride ion may play a role in the mechanism of action of sulfuryl fluoride in insects and also in mammals. Serum fluoride levels were elevated over control values in several species of laboratory ­ animals following acute or subchronic exposure to sulfuryl fluoride. Some of the observations in the animals at higher concentrations of sulfuryl fluoride were consistent with acute fluoride poisoning (Drill, 1954; Goodman et al., 1980; Greenwood, 1940; Nitschke et al., 1986; Whitford, 1996). The pharmacokinetics and metabolism of inhaled sulfuryl fluoride were evaluated in male rats exposed to 30 or 300 ppm 35 S-labeled sulfuryl fluoride for 4 h (Mendrala et al., 2005). Sulfuryl fluoride was rapidly absorbed and achieved maximum concentrations of radioactivity in the plasma near the end of the 4-h exposure. Radioactivity rapidly cleared from the plasma (initial half-life of 2.5 h after 30-ppm and 1–2.5 h after 300-ppm exposures) and was rapidly excreted as fluorosulfate (FSO3) and sulfate primarily in the urine. Seven days postexposure, small amounts of radioactivity were distributed among several tissues, with the highest concentration in the lungs and nasal tissues, which is indicative of the primary role of the upper and lower respiratory tract in the absorption and hydrolysis of inhaled sulfuryl fluoride. Radiochemical profiles did not provide any evidence of parent 35S-labeled sulfuryl fluoride in the blood. Identification of FSO3 and sulfate in blood and urine suggests that sulfuryl fluoride is hydrolyzed to fluorosulfate, with release of fluoride, followed by further hydrolysis to sulfate and release of the remaining fluoride. This metabolism is supported by the dose-dependent increases in fluoride detected in the blood, tissues, and urine following exposure of rats to 30 or 300 ppm nonradiolabeled sulfuryl fluoride. Rats exposed to 30 and 300 ppm sulfuryl

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fluoride had urinary concentrations of fluoride three and eight times higher than control animals, respectively. Mendrala et al. (2005) indicated that summation of 35 S-sulfuryl fluoride-derived radioactivity recovered in the urine, feces, and tissues would give an approximation of the total absorbed radioactivity following single 4-h exposures of rats to 30 or 300 ppm sulfuryl fluoride. The total absorbed radioactivity could be used as an estimate of the amount of sulfuryl fluoride that was absorbed. Approximately 688 g-equivalent sulfuryl fluoride (6.7 mol-equivalent) were absorbed during a 4-h 30-ppm exposure and approximately 85% was eliminated in the urine. Approximately 5690 g-Eq sulfuryl fluoride (55 mol-equivalent) were absorbed during a 4-h 300-ppm exposure and 81% was eliminated in the urine. The total internal dose based on radioactivity was compared to a calculated amount of sulfuryl fluoride inhaled during the 4-h exposure based on the body weight of these rats and a standard inhalation rate (Blackburn, 1988). These comparisons indicated that rats exposed to 30 ppm absorbed 14.0% of the inhaled sulfuryl fluoride, while rats exposed to 300 ppm absorbed 12.5% of the inhaled sulfuryl fluoride.

103.5  Toxicology in humans There have been a few medical case reports of individuals exposed to sulfuryl fluoride. A 30-year-old male was exposed to unknown concentrations of sulfuryl fluoride in air containing 1% chloropicrin for 4 h (Taxay, 1966). Nausea, vomiting, cramps, abdominal pain, and itching were reported during exposure to sulfuryl fluoride. Vital signs were normal upon admittance to the hospital; however, reddening of the conjunctival, pharyngeal, and nasal mucosae, diffuse rhonchi, and paresthesia of the lateral surface of the right leg were reported. Serum fluoride levels were elevated above normal values. The signs and symptoms resolved quickly and the patient was discharged from the hospital after 4 days. In another case report, an elderly couple returned to their home approximately 5–8 h after their house was ventilated following approximately 24-h fumigation of their home (Nuckolls et al., 1987). Within 24 h of their return, the wife experienced weakness, nausea, and repeated vomiting and her husband complained of dyspnea and restlessness. Within 48 h, the husband had a generalized seizure followed by cardiopulmonary arrest. The wife died within 7 days due to ventricular fibrillation. Serum fluoride level of the wife 6 days after the house was fumigated was 0.5 mg/l (background levels are highly dependent upon fluoride levels in drinking water and range from 0.010 to 0 .2 mg/l; Burtis and Ashwood, 1999). Individuals have entered structures during fumigation with sulfuryl fluoride and were found dead or died shortly after exposure (Scheuerman, 1986). The cause of death appeared to be severe pulmonary edema with congestion. Serum fluoride levels in two individuals were elevated above normal values. A more recent case report (Schneir

Hayes’ Handbook of Pesticide Toxicology

et al., 2008) describes an individual removed from the tarpaulin used to enclose an apartment complex being fumigated with sulfuryl fluoride. The patient displayed local irritant effects from the warning agent, chloropicrin, as well as systemic effects of sulfuryl fluoride. Blood fluoride concentration was 24 mg/l. The authors indicated that the clinical course included initial hypocalcemia, delayed onset of ventricular dysrhythmias, and death, and that these effects were consistent with fluoride being the predominant mechanism of toxicity. Groups of fumigators using sulfuryl fluoride and/or methyl bromide were evaluated in a neurobehavioral battery (Anger et al., 1986). There were no statistically significant differences over approximately 70 endpoints for the fumigators exposed to sulfuryl fluoride as compared to a reference group. The sulfuryl fluoride group had more symptompositive reports for the lower extremities than the referents; however, the sulfuryl fluoride group performed better than the referent group on all three tests of tactile sensitivity. A statistically nonsignificant reduced performance in all cognitive tests was found in the sulfuryl fluoride group compared to the control group in the presence of a nonsignificant increase of illegal drug use and of drinks per week, as well as a decrease in educational level in the sulfuryl fluoride group compared to the control group. Thus, the slight difference in cognitive test results was likely due to differences other than exposure to sulfuryl fluoride. A cross-sectional study of structural fumigation workers included evaluation of nerve conduction, vibration, neurobehavioral, visual, olfactory, and renal functional testing (Calvert et al., 1998). Exposure to methyl bromide or sulfuryl fluoride was defined as years employed and the percent of jobs in the past year using methyl bromide or sulfuryl fluoride. Sulfuryl fluoride exposure over the year preceding testing was associated with reduced performance on the Pattern Memory Test and olfactory testing. However, other endpoints related to memory were all negative for an association and the authors state that the pattern of memory findings may have arisen by chance. Fumigation workers performed worse on tests of median nerve function than did the referents and the authors attributed this finding to ergonomic stresses of the job. The authors reported that a 1991 NIOSH study found that more than two-thirds of actual fumigant-worker exposures to sulfuryl fluoride, as measured with personal airborne sampling, were undetectable and that all were below the Occupational Safety and Health Administration permissible exposure limits.

103.6  Conclusion 103.6.1  Risk Characterization Remarkably little difference was observed among laboratory animal species exposed to sulfuryl fluoride in studies that varied from 13 weeks to 2 years in duration. The

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Table 103.4  Summary of NOELs for Sulfuryl Fluoride Toxicity Studies Repeated exposure studies

General toxicity NOEL (ppm)

Neurotoxicity NOEL (ppm)

2-Week rat

100

600

2-Week rabbit

100

100

2-Week mouse

30

30

2-Week dog

100

100

13-Week rat

30

100

13-Week rat (electrophysiology)

30

30

13-Week rabbit

30

100 (M), 30 (F)

13-Week mouse

30

30

13-Week dog

100

100

12-Month rat

5 (M)a, 20 (F)

80

12-Month rat (Neurotox guideline)

80

80

12-Month mouse

20

20

12-Month dog

20

80 a

24-Month rat

5 (M) , 20 (F)

80

18-Month mouse

20

20

Two-generation rat reproduction

5

20

a

Microscopic dental fluorosis was observed in several 20-ppm male rats.

NOEL for general toxicity in the regulatory guideline studies of up to 2 years in duration varied from 100 to 5 ppm (Table 103.4). The NOEL of 5 ppm for male rats on the 2-year study was based on very sight dental fluorosis in several male rats at 20 ppm; since dental fluorosis is not considered an adverse effect, the more relevant value for risk assessment would be an NOAEL of 20 ppm for the 2-year rat study. The NOEL of 5 ppm in the two-generation rat study was based on minimal alveolar histiocytosis in the lungs of parental rats at 20 ppm. This portal-of-entry irritative effect in the lungs was attributed to the 6-h/day, 7-days/week exposures during mating, gestation, and lactation for the reproduction study as compared to the 6-h/day, 5-days/week exposures on the other rat studies, which did not result in alveolar histiocytosis at 20 ppm, even after 2 years of repeated exposure. Thus, the lower NOEL of 5 ppm for the irritative effects in the lungs of the adult rats on the reproduction study would not be relevant for standard risk assessments because of the 7-days/week exposures and the fact that the minimal changes in the lungs are portal-of-entry effects rather than systemic toxicity. In general, NOELs for neurotoxicity were higher than the NOELs for general toxicity.

The label instructions specify sealing structures to confine the gas during the fumigation with tarps and/or sealing (Dow AgroSciences, 2005, 2006). After the appropriate fumigation period, the building is aerated using specified aeration procedures. Depending upon which aeration procedure is used, the building must be secured for either 6 or 8 h. After this waiting period, the concentration of sulfuryl fluoride in the breathing zones must be determined. The structure must be ventilated until the concentration of sulfuryl fluoride is equal to or less than 1 ppm, at which point the structure may be reoccupied. The fumigation site cannot be reoccupied until aeration is complete. Only an approved detection device of sufficient sensitivity, such as the INTERSCAN, MIRAN [SapphIRe], or Spectros ExplorIR gas analyzer, can be used to confirm a concentration of sulfuryl fluoride is 1 ppm or less. Warning signs must remain posted until aeration is determined to be complete.

References ACGIH. (1998). TLVs and other Occupational Exposure Values - 1998. Albee, R.R., Spencer, P.J., and Bradley, G.J. (1993). “Sulfuryl Fluoride: Electrodiagnostic, FOB, and Motor Activity Evaluation of Nervous System Effects from Short-term Exposure”. Unpublished report of The Dow Chemical Company. Anger, W. K., Moody, L., Burg, J., Brightwell, W. S., Taylor, B. J., Russo, J. M., Dickerson, N., Setzer, J. V., Johnson, B. L., and Hicks, K. (1986). Neurobehavioral evaluation of soil and structural fumigators using methyl bromide and sulfuryl fluoride. Neurotoxicology 7, 137–156. Blackburn, K. (1988). “Recommendations for and Documentation of Biological Values for Use in Risk Assessment,” EPA/600/6-87/008. ORD, U.S. EPA, Cincinnati, Ohio. Boorman, G. A., Eustis, S. L., Elwell, M. R., Montgomery, C. A. Jr., and MacKenzie, W. F. (1990). “Pathology of the Fischer Rat”. Reference and Atlas. Academic Press, Inc, San Diego, CA. Bradley, G. J., Landry, T. D., Battjes, J. E., and Quast, J. F. (1990). “Sulfuryl Fluoride: Four-hour Dermal Vapor Exposure in Fischer 344 Rats”. Unpublished report of The Dow Chemical Company. Breslin, W. J., Liberacki, A. B., Kirk, H. D., Bradley, G. J., and Crissman, J. W. (1993). Sulfuryl fluoride: Two-generation inhalation reproduction study in Sprague-Dawley rats. Toxicologist 13, 368. Burtis, C. A., and Ashwood, E. R. (1999). “Tietz Textbook of Clinical Chemistry,” 3rd ed. W.B. Saunders Co, Philadelphia, PA. Cady, G. H., and Misra, S. (1974). Hydrolysis of sulfuryl fluoride. Inorganic Chem. 13, 841–873. Calvert, G. M., Mueller, C. A., Fajen, J. M., Chrislip, D. W., Russo, J., Briggle, T., Fleming, L. E., Suruda, A. J., and Steenland, K. (1998). Health effects associated with sulfuryl fluoride and methyl bromide exposure among structural fumigation workers. Amer. J. Public Health 88, 1774–1780. Dow AgroSciences. (1997). General information on Vikane Gas Fumigant; Product Brochure #311-56-077. Dow AgroSciences. (2005). ProFume Gas Fumigant Specimen Label, revised 7-19-05. Dow AgroSciences. (2006). Vikane Specialty Gas Fumigant Label, revised 5-31-06.

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Drill, V. A. (1954). “Pharmacology in Medicine”. McGraw-Hill, New York. Eisenbrandt, D. L., and Nitschke, K. D. (1989). Inhalation toxicity of sulfuryl fluoride in rats and rabbits. Fundam. Appl. Toxicol. 12, 540–557. Eisenbrandt, D. L., Williams, D. M., Albee, R. R., and Streeter, C. M. (1987). “Sulfuryl Fluoride (Vikane Gas Fumigant): An Ultrastructural Assessment of the Lungs of Rats Exposed to High Concentrations of Sulfuryl Fluoride.” Unpublished report of The Dow Chemical Company. Gollapudi, B. B., McClintock, M. L., and Nitschke, K. D. (1990a). “Evaluation of Sulfuryl Fluoride in the Mouse Bone Marrow Micronucleus Test.” Unpublished report of The Dow Chemical Company. Gollapudi, B. B., Samson, Y. E., and Zempel, J. A. (1990b). “Evaluation of Sulfuryl Fluoride in the Ames Salmonella/Mammalian Microsome Bacterial Mutagenicity Assay.” Unpublished report of The Dow Chemical Company. Gollapudi, B. B., McClintock, M. L., and Zempel, J. A. (1991). “Evaluation of Sulfuryl Fluoride in the Rat Hepatocyte Unscheduled DNA Synthesis (UDS) Assay.” Unpublished report of The Dow Chemical Company. Goodman, A. G., Goodman, L. S., and Gilman, A. (1980). “The Pharmacological Basis of Therapeutics,” 6th ed. Macmillan, New York. Gorzinski, S. J., and Streeter, C. M. (1985). “Effect of Acute Vikane Exposure on Selected Physiological Parameters in Rats.” Unpublished report of The Dow Chemical Company. Greenwood, D. A. (1940). Fluoride intoxication. Physiol. Rev. 20, 582–616. Hanley, T. R. Jr., Calhoun, L. L., Kociba, R. J., and Greene, J. A. (1989). The effects of inhalation exposure to sulfuryl fluoride on fetal development in rats and rabbits. Fundam. Appl. Toxicol. 13, 79–86. Jones, M. M., and Lockhart, W. L. (1968). Kinetics of decomposition of the fluorosulphate ion in aqueous solution. J. Inorg. Nucl. Chem. 30, 1237–1243. Landry, T. D., and Streeter, C. M. (1983). “Sulfuryl Fluoride: Effects of Acute Exposure on Respiration in Rats.” Unpublished report of The Dow Chemical Company. Lockwood, D. T. (1958). “Results of Dietary Feeding of Rats with Feed Fumigated with Sulfuryl Fluoride (Vikane).” Unpublished report of The Dow Chemical Company. Mattsson, J. L., Albee, R. R., Eisenbrandt, D. L., and Chang, L. W. (1988). Subchronic neurotoxicity in rats of the structural fumigant, sulfuryl fluoride. Neurotoxicol. Teratol. 10, 127–133. Meikle, R. W., Steward, D., and Globus, O. A. (1963). Fumigant mode of action, drywood termite metabolism of Vikane fumigant shown by labelled pool technique. J. Agr. Food Chem. 11, 226–230. Mendrala, A. L., Markham, D. A., and Eisenbrandt, D. L. (2005). Rapid uptake, metabolism, and elimination of inhaled sulfuryl fluoride fumigant by rats. Toxicol. Sci. 86, 239–247. Miller, R. R., Calhoun, L. L., Keyes, D. G., and Kociba, R. J. (1980). “Sulfuryl Fluoride (Vikane Fumigant): An LC50 Determination.” Unpublished report of The Dow Chemical Company. Mohr, U., Dungworth, D. L., and Capen, C. C. (1992). “Pathobiology of the Aging Rat.” International Life Sciences Institute, Washington, D.C. Nitschke, K. D., and Lomax, L. G. (1989). “Sulfuryl Fluoride: Acute LC50 Study with B6C3F1 Mice.” Unpublished report of The Dow Chemical Company. Nitschke, K. D., and Quast, J. F. (1990). “Sulfuryl Fluoride: Acute LC50 Study with CD-1 Mice.” Unpublished report of The Dow Chemical Company.

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Nitschke, K. D., and Quast, J. F. (1991). “Sulfuryl Fluoride: Two-week Inhalation Toxicity Study in Beagle Dogs.” Unpublished report of The Dow Chemical Company. Nitschke, K. D., and Quast, J. F. (1993). “Sulfuryl Fluoride: Thirteen-week Inhalation Toxicity Study in CD-1 Mice.” Unpublished report of The Dow Chemical Company. Nitschke, K. D., and Quast, J. F. (1995). “Sulfuryl Fluoride: Two-week Inhalation Toxicity Study in CD-1 Mice.” Unpublished report of The Dow Chemical Company. Nitschke, K. D., Albee, R. R., Mattsson, J. L., and Miller, R. R. (1986). Incapacitation and treatment of rats exposed to a lethal dose of sulfuryl fluoride. Fundam. Appl. Toxicol. 7, 664–670. Nitschke, K. D., Beekman, M. J., and Quast, J. F. (1992). “Sulfuryl Fluoride: 13-week Inhalation Toxicity Study in Beagle Dogs.” Unpublished report of The Dow Chemical Company. Nuckolls, J. G., Smith, D. C., Walls, W. E., Oxley, D. W., Hackler, R. L., Tripathi, R. K., Armstron, C. W., Miller, G. B. (1987). Fatalities resulting from sulfuryl fluoride exposure after home fumigation Virginia. JAMA 258 15;2041–2044. Quast, J. F., Beekman, M. J., and Nitschke, K. D. (1993a). “Sulfuryl Fluoride: One-year Inhalation Toxicity Study in Beagle Dogs.” Unpublished report of The Dow Chemical Company. Quast, J. F., Bradley, G. J., and Nitschke, K. D. (1993b). “Sulfuryl Fluoride: 18-month Inhalation Oncogenicity Study in CD-1 Mice.” Unpublished report of The Dow Chemical Company. Quast, J. F., Bradley, G. J., and Nitschke, K. D. (1993c). “Sulfuryl Fluoride: 2-year Inhalation Chronic Toxicity/Oncogenicity Study in Fischer 344 Rats.” Unpublished report of The Dow Chemical Company. Scheuerman, E. H. (1986). Suicide by exposure to sulfuryl fluoride. J. Forensic Sci. 31, 1154–1158. Schneir, A., Clark, R. F., Kene, M., and Betten, D. (2008). Systemic fluoride poisoning and death from inhalational exposure to sulfuryl fluoride. Clin. Toxicol. (Phil.) 46, 850–854. Spencer, P. J., Bradley, G. J., and Quast, J. F. (1994). “Sulfuryl Fluoride: Chronic Neurotoxicity Study in Fischer 344 Rats - Final Report.” Unpublished report of The Dow Chemical Company. Taxay, E. P. (1966). Vikane inhalation. J. Occup. Med. 8, 425–426. Torkelson, T. R. (1959) “Summary Report of Toxicological Studies with Vikane (Sulfuryl Fluoride, SO2F2).” Unpublished report of The Dow Chemical Company. Torkelson, T. R., Hoyle, H. R., and Rowe, V. K. (1966). Toxicological hazards and properties of commonly used space, structural and certain other fumigants. Pest Control July, 1–8. U.S. Environmental Protection Agency (1982). “Pesticide Assessment Guidelines, Subdivision F, Hazard Evaluation: Human and Domestic Animals,” pp. 98–100. U.S. Environmental Protection Agency, Washington, DC. Vernot, E. H., MacEwen, J. D., Haun, C. C., and Kinkead, E. R. (1977). Acute toxicity and skin corrosion data for some organic and inorganic compounds and aqueous solutions. Toxicol. Appl. Pharmacol. 42, 417–423. Whitford, G. (1996). The metabolism and toxicity of fluoride. In “Monographs in Oral Science” (H. Myers, ed.), 2nd ed., Vol. 16. Karger AG, Basel.

Chapter 104

Phosphine A. V. Lyubimov1 and V. F. Garry2 1

University of Illinois at Chicago University of Minnesota

2

104.1  Identity, properties, and uses Chemical Name: Hydrogen Phosphide Structure:  PH3. Synonyms: Phosphoretted hydrogen, phosphorus hydride, Phosphorus trihydride. The CAS Registry No.: 7803-51-2. Conversion factor: 1 ppm  1.39 mg/m3. The most common commercial fumigants generating phosphine are aluminum phosphide and magnesium phosphide. Aluminum Phosphide (AIP) is sold under the following trade names: Phostoxin, Fumitoxin, Agtoxin, Weevilcide, Gastoxin, Phosfume, Fastphos. Common trade names for Magnesium Phosphide (Mg3P2) are Fumi-Cel, Fumi Strip, Magtoxin, Magnaphos, Magphos.

104.1.1  Physical Properties Pure phosphine is an odorless and colorless gas with a molecular weight of 34.00 and density of 1.17 at 25°C. Commercial grade phosphine derived from aluminum or magnesium phosphide can contain to a variable degree higher molecular weight phosphines including diphosphines. These higher phosphines give commercial grade fumigants containing aluminum or magnesium phosphide odor characteristics described as decaying fish or “garlic-like.” Commercial grade phosphine containing diphosphines can ignite and form explosive mixtures at concentrations exceeding 1.8% phosphine in air. The rate of conversion of the phosphide to phosphine is temperature and humidity dependent. Similarly, metal phosphides readily hydrolyze in water to yield phosphine, which is poorly soluble in water. Major products resulting from the

Hayes’ Handbook of Pesticide Toxicology Copyright © 2001 Elsevier Inc. All rights reserved

o­ xidation of phosphine in water are hypophosphorous and phosphoric acids (Van Wazer, 1958; WHO, 1988).

104.1.2  Chemistry Phosphine is a nucleophile and acts as a strong reducing agent (Lam et al., 1991). Under standard conditions of temperature, pressure, and humidity PH3 is stable and does not undergo autoxidation. Very early work suggests that under conditions of increased atmospheric pressure and oxygen content autoxidation can occur (Van Wazer, 1958). Further, in the presence of trace levels of diphosphine and perhaps other higher phosphines in air, PH3 will undergo a branched chain oxidation reaction (Green et al., 1984; Osadchenko and Tomilov, 1969), a form of autooxidation. Similarly, under experimental conditions the reaction can be induced photolytically by ultraviolet (UV) light or ammonia (Buchanan and Hanrahan, 1970; Woller, 1965). The branched chain reaction when it occurs is a generator and a good source of free radicals (see below) (Green et al., 1984): Propagation O2  PH2•  ⇒  HPO  OH• OH•  PH3  ⇒  PH2•  H2O O2  PH2•  ⇒  PH  HO2• O2  PH  ⇒  HPO  O Branching O  PH3  ⇒  PH2•  OH• Termination O  O2  M  ⇒  O3  M Radical  Wall  ⇒  Compound Secondary Reactions HPO  O2  ⇒  HPO3 HPO3  H2O  ⇒  H3PO4

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104.2  Sources, uses, and formulations

temperature, and/or ammonia as well. The studies reviewed below emerge as a complex picture of the toxicant effects of phosphine.

104.2.1  Natural Sources

104.2.2  Commercial Sources Metal phosphides, notably aluminum, magnesium, and zinc phosphide, are the most common commercial sources of phosphine. Aluminum and magnesium phosphides are commonly used fumigants supplied as pellets, tablets, sachets, ropes, or strips (Meister, 1999) for insect control in stored grains and other products. Zinc phosphide baits are commonly used for rodent control. Ammonia from ammonium carbamate is sometimes used as a warning odorant in some fumigant formulations. Phosphine gas is also used in the synthesis of flame retardants, as a dopant in the semiconductor industry, and as a polymerization initiator and catalyst (U.S. DHHS, 1993).

104.3  Toxicology 104.3.1  Overview The modern history of our understanding of the biologic effects of the toxicant phosphine begins with the works of O. R. Klimmer (1969, 1970). In these works the investigator established the dose related lethal effects of phosphine in multiple species, determined the dose threshold for lethality, and explored possible mechanisms for lethality including effects on hemoglobin. From these works, there is ample evidence that the acute lethal effects of phosphine can occur at levels less than 8 mg/m3. Since that time, work by others has gone forward to explore the avenues for the lethal effects of this toxicant gas in insects, mammals, and humans in vivo and in vitro. Genotoxicity and reproductive effects have also been considered. Because phosphine is an explosive hazard, many of the laboratory-based studies have been conducted under exposure conditions to eliminate or reduce the possibility of the branched chain oxidation reaction in air. Thus, these studies reflect the effects of phosphine in the unoxidized state. Human case and field population studies and some in vitro studies may reflect to a greater or lesser degree the toxicant effects of phosphine and its auto-oxidation products induced by the contaminant diphosphine in the commercial product, and uncontrolled environmental conditions including UV light, humidity,

104.4  Toxicity and mode of action 104.4.1  Acute Toxicity 104.4.1.1  Symptoms Early on, Klimmer found that animals exposed to high concentrations of phosphine quickly develop lassitude, ataxia, apnea, and cardiovascular collapse resulting in death within one half-hour (Klimmer, 1969). At lower concentrations (range studied 7.5 to 564 mg/m3) time to death varied with dose (Fig. 104.1). Concentrations as low as 7 mg/m3 are lethal over a period of 820 hours. Toxicity to humans occurs either due to ingestion of aluminium phosphide or inhalation of phosphine (PH3), liberated from fumigated grains. In humans, case studies involving suicide and suicide attempts by ingestion of pellets of aluminum phosphide are instructive. Poisoning due to ingestion of aluminium phosphide is still common in the Indian subcontinent (Kaushik et al., 2007) and represent the most important cause of fatal chemical toxicity in Iran (Mehrpour et al., 2008). After ingestion of aluminum phosphide, phosphine gas is released in the stomach and is rapidly absorbed throughout the gastrointestinal tract (Chugh, 2003; Olson, 2006). Case studies and case reports have shown that exposure to phosphine gas is an important cause of morbidity, especially in those workers who apply either aluminum phosphide

1000 500 Concentration (mg/m3)

Phosphine can be generated in decaying organic matter in open air sewage treatment plants (Dévai et al., 1998) and other sources of decaying organic material (Glindemann et al., 1996) including landfills, compost processing, and river sediments. Maximum concentrations detected were approximately 20 ppb.

200 100 50 20 10 5 2 1 0.1

1

10 Time (hrs)

100

1000

Figure 104.1  Comparison of dose and time to death in different species. Response of rats (), rabbit (), guinea pigs (•) and cats () to phosphine. Each point indicates the concentration of phosphine to which a group of animals was exposed and the average time to death. From data of Klimmer (1969) and Hayes and Laws (1991). Reproduced with permission.

Chapter | 104  Phosphine

pellets as fumigant or, more commonly, those workers who entered containers of stored grain treated with aluminum phosphide and do not directly handle the fumigant (Kurzbauer and Kiesler, 1987; Burgess et al., 2000). Brautbar and Howard (2002) reviewed the toxic effects of phosphine gas on the lungs and central nervous system in two workers. Sudakin (2005) presented a case report of suspected inhalation exposure to phosphine gas in a manufacturing facility for aluminium phosphide fumigants, which was associated with acute dyspnea, hypotension, bradycardia and other signs of intoxication. Interestingly, the case reports of inhalation exposure to phosphine gas described in non-fumigant-related exposures occurred in association with the clandestine synthesis of methamphetamine, as phosphine can be produced from phosphoric acid that is formed when iodine and red phosphorus are combined in aqueous media (WillersRusso, 1999; Burgess, 2001). Most recently Kaushik et al. (2007) reported a case of aluminium phosphide poisoning with systemic toxicity in a 16-year-old patient who developed a previously undocumented complication of subendocardial infarction, with characteristic electrocardiographic changes reverting back to normal after a period of 10 weeks, even though the patient had clinical recovery much earlier. Rapid onset of epigastric distress, hypotension, cardiovascular collapse, and death are a recurrent pattern. In those who reach a hospital, altered sensoria, vomiting, severe acidosis, hypotension, cardiac arrhythmia, jaundice, and pulmonary crepitation were common occurrences (Banjaj and Wasir, 1988; Misra et al., 1988a, b; Singh et al., 1996). There are no well-established biomarkers of phosphine poisoning. Mephrpour et al. (2008) found a correlation of hyperglycemic effect and mortality in the prospective study conducted across a 14-month period in 45 patients. In a review of 195 intentional intoxication cases, Singh et al. (1985) concluded that ingestion of 1.5 g aluminum phosphide can be lethal in adults. Other cases of fatalities as a result of accidental exposure, suicide, and even possible homicide, have also been noted (Wilson et al., 1980; Siwach et al., 1995; Singh et al., 1996; NIOSH Alert, 1999; Abner-Rahman, 1999, 2000; Bogle et al., 2006; Shadnia, et al., 2008). Most fatalities due to phosphine toxi­city occur within 24 hours of exposure. Autopsy findings from published accidental death investigations (Garry et al., 1993; Heyndrickx et al., 1976; Wilson et al., 1980) show microscopic pulmonary congestion with edema and alveolar cell necrosis, individual myocardial cell and liver cell necrosis, and anoxic changes in the brain. Klimmer (1970) noted earlier in autopsied animals a peculiar ­crimson color to the blood. These findings were variably recorded in the human autopsy and in clinical case ­studies. These human case studies and early animal acute ­toxicity studies provide some insights for formal mechanistic studies.

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104.5  Animal dose/response 104.5.1  Threshold for Lethality The early studies of Klimmer (1969) as illustrated previously show that for rats, rabbits, cats, and guinea pigs, a threshold for acute lethality by the inhalation route occurs at about 7 mg/m3. Similarly, Newton et al. (1993), demonstrated in pregnant Fischer 344 female rats, exposed six hours daily, four days was the median lethal time at a concentration of 9.7 mg/m3. Concentrations below 7 ppm showed no lethality. In mice (both sexes) the Median Lethal Dose after two weeks exposure is 9 mg/m3 (Barbosa et al., 1994). Concentrations below this level were not lethal. As indicated before, there are only minor differences in the mortality data from earlier to more recent studies regarding duration time-dose threshold for acute lethality.

104.5.2  Acute and Subacute   Dose/Response Given the time-duration effects noted above and other factors, the LC50 for inhaled phosphine is somewhat variable. Early studies by Waritz and Brown (1975) showed a four hour LC50 of 11 ppm in male rats. Using highly purified phosphine, Omae et al. (1996) reported a four hour LC50 between 26.5 and 33.4 ppm in male mice. Newton et al. (1993) reported no lethality in male and female rats acutely exposed to 10 ppm phosphine for six hours. In subacute studies in Fischer 344 female rats these authors indicate that three day exposure to 10 ppm phosphine was lethal. They further demonstrated that female rats were more sensitive to the lethal effects of the inhaled gas.

104.6  Absorption, distribution, metabolism, and excretion Aside from empirical observations regarding ingestion and respiratory exposure, there is little toxicokinetic data regarding absorption, distribution, and excretion of phosphine and its reaction products. In one study, 32P labeled phosphine as reaction product residues (hypophosphite and phosphite) in flour were fed to mice. Labeled material in excreta was found to persist for periods up to three weeks (Robinson and Bond, 1970).

104.7  Cellular and molecular studies 104.7.1  General Much of the work regarding phosphine as a metabolic poison centers on the concept that reactivity of phosphine as a nucleophile, and/or the electrophilic character of the

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i­ntermediates arising from oxidation, could lead to derivatization of critical biomolecules (Lam et al., 1991). Certain critical biologic endpoints including the cytochromes and cytochrome oxidase system, hemoglobin, peroxidases and lipid peroxidation, catalase, cholinesterase, and DNA have been studied in some detail. The reported phosphine effects in each of these systems will be discussed below. As part of the discussion, the importance of oxygen as a modifier enhancing toxicity and the reduction of toxicity in reduced oxygen atmospheres will be considered.

104.7.2  Cytochromes and Cytochrome Oxidase Early on, the requirement for oxygen to mediate the toxicity of phosphine was identified in insects (Bond et al., 1969) indicating that the gas may be an aerobic mitochondrial respiratory poison. Since that time, in vitro studies, both animal and insect, have shown that the respiratory enzyme, cytochrome c oxidase, may be the specific site of action (Bolter and Chefurka, 1990; Chaudhry, 1997; Kashi and Chefurka, 1976; Price, 1980). On the other hand, in vivo treatment of insects with lethal dose levels of phosphine (Nakakita, 1987) showed no more than 50% inhibition of the enzyme. Further work showed that this level of respiratory enzyme inhibition was sufficient to generate superoxide anions (Bolter and Chefurka, 1990) and these authors suggested that the toxicity of phosphine was due to free radical damage. Toxicological studies in mammals have shown that glutathione (GSH) provides important protection against phosphineinduced disruptions to cells as GSH levels were found to decrease in rat tissue and human blood following phosphine exposure (Chugh, 1997, Hsu et al., 2000, 2002a, 2002b). Conversely, addition of GSH to mouse cells partially protected them against phosphine-induced cell death, reactive oxygen species, and DNA damage (Hsu et al., 1998). The GSH depletor buthionine sulfoximine, which irreversibly inhibits c-glutamylcysteine synthetase at the first step of GSH synthesis, can further lower the already reduced GSH levels in rats exposed to phosphine (Hsu et al., 2002b), although the effect on mortality of cotreatment with phosphine and buthionine sulfoximine was not determined. Furthermore, GSH depletion is reported to have no effect on phosphine-induced mortality in insects (Chaudhry and Price, 1992). Cha’on et al. (2007) reported the following dosedependent actions of phosphine, in vitro: (1) reduction of ferric iron (Fe3) to ferrous iron (Fe2), (2) release of iron from horse ferritin, (3) and the peroxidation of lipid as a result of iron release from ferritin. Using in situ hybridization, they showed that the ferritin genes of C. elegans, both ferritin-1 and ferritin-2, are expressed along the digestive tract with greatest expression at the

Hayes’ Handbook of Pesticide Toxicology

proximal and distal ends. Basal expression of the ferritin-2 gene, as determined by quantitative PCR, is approximately 80 times that of ferritin-1. However, in response to phosphine, transcript levels of ferritin-1 are induced at least 20-fold higher levels, whereas there is no change in the level of ferritin-2. This resembles the reported pattern of ferritin gene regulation by iron, suggesting that phosphine toxicity may be related to an increase in the level of free iron. Indeed, iron overload increases phosphine toxicity in C. elegans at least threefold. Moreover, the authors demonstrated that suppression of ferritin-2 gene expression by RNAi (RNA interference assay) significantly increases sensitivity to phosphine. Zuryn et al. (2008) employed Caenorhabditis elegans to investigate the effects of phosphine on its proposed in vivo target, the mitochondrion. Authors found that phosphine rapidly perturbs mitochondrial morphology, inhibits oxidative respiration by 70%, and causes a severe drop in mitochondrial membrane potential within 5 h of exposure

104.7.3  Hemoglobin In seminal efforts Trimborn and Klimmer (1962) described phosphine-induced hemoglobin denaturation, oxidation to methemoglobin, and formation of a peculiar pigmented form of hemoglobin “Verdichromogen.” Studies of purified hemoglobin by Potter et al. (1991) and Chin et al. (1992) showed that with increasing duration of exposure, phosphine in concentrations as low as 0.11 M gradually resulted in the formation of hemichrome pigment. In intact red blood cells, Potter et al. (1991) noted formation of Heinz bodies (hemoglobin protein aggregates) at PH3 concentrations as low as 2 g/ml. The toxicant effects both in intact cells and in purified hemoglobin were abolished by incubation in a reduced oxygen atmosphere, indicating an oxygen requirement for phosphine hemoglobin toxicity.

104.8  Peroxidases, lipid peroxidation, catalase, and cholinesterase Because phosphine is a strong reducing agent, peroxidation and formation of peroxides and their reduction are concerns mechanistically and therapeutically. Studies by Pazynich et al. (1984) in animals showed that the gas inhibited myeloperoxidase enzyme at concentrations of 8 mg/m3. In the occupational setting, Garry et al. (1990) noted histochemically a 50% reduction in myeloperoxidase activity in neutrophils from exposed workers compared to control subjects. Ambient air monitoring data obtained at the time varied from 0.4 to 5.8 mg/m3. The permissible exposure limit (PEL) for phosphine at the time in the US was 0.4 mg/m3. In other human studies by Chugh et al. (1996) in 45 patients recovering from phosphine

Chapter | 104  Phosphine

poisoning, serial studies of serum levels of superoxide dismutase (SOD), malondialdehyde (MDA), and catalase were performed. Increased levels of SOD and MDA were found in nonsurvivors while catalase was inhibited. Remarkably similar findings (i.e., decreased peroxidase and catalase and increased superoxide dismutase) were reported by Bolter and Chefurka (1990) and Chaudhry and Price (1990) in insects. Taken together, the works cited above indicate that phosphine intoxication can lead to accumulation of cellular peroxides. Further, the oxidation of phosphine (Lam et al., 1991) can lead to formation of reactive phosphorylating species. As such effects on cholinesterase are also possible. Significant inhibition of cholinesterase was detected in animals (Pazynich et al., 1984). Occupational studies of grain fumigant applicators (Potter et al., 1993) and in vitro studies in human red blood cells (Potter et al., 1991) demonstrate that significant phosphine-induced inhibition of red cell cholinesterase occurs at concentrations exceeding 10 g/ml.

104.9  Genotoxicity, cancer and reproductive effects Studies in the occupational setting (Garry et al., 1989, 1990, 1992) suggest that in enclosed space applications where PH3 ambient air concentrations exceed the ­permissible exposure limit of 0.4 mg/m3 (range 0.4–5.8 mg/m3) for a duration of more than 20 minutes, increased chromosome aberrations are detectable in human lymphocytes from exposed workers. Studies by Barbosa and Bonin (1994) using micronucleus assay found no increase in micronucleus frequency in exposed workers where ambient exposures were less than the PEL. Later follow up studies by Garry et al. (1996) of the same worker population did not show increased chromosome aberrations. During the interim, changes in application practice from manual probe application to more automated methods and nonuse of phosphine in pesticide applications were noted (unpublished). In subacute tightly controlled animal studies (Kligerman et al., 1994b) using purified PH3 mixed with nitrogen, no increased numbers of micronuclei or chromosome aberrations were found in spleen cells cultured from animals exposed to phosphine for six hours per day for 9 days at concentrations as high as 7 mg/m3 in ambient air. A single 6 hours 20 mg/m3 study by this investigator showed similar negative results (Kligerman et al., 1994a). In similarly constructed subchronic studies, Barbosa et al. (1994) found significantly increased numbers of micronuclei at the highest concentration tested (6.3 g/ m3). Cast in the light of these in vivo studies (both animal and human) one can conclude that in regard to genotoxicity, phosphine may be a genotoxin. From a mechanistic view, in vitro studies of genotoxicity offer some additional insights. In these studies (Garry et al., 1989; Hsu et al., 1998) aluminum or magnesium phosphide was used as a phosphine

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generating system. Exposure of human lymphocytes (Garry et al., 1989) to concentrations of phosphine (1.4–4.5 g/l) derived from AlP for 20 minutes yielded increased chromosome aberrations after 96 hours of lymphocyte culture, indicating that the expression of genotoxicity of phosphine is delayed. In a much more detailed mechanistic examination of the genotoxicity of phosphine derived from AlP or Mg3P2 in Hepa cells at a nominal concentration of 1 mM PH3, Hsu et al. (1998) found that reactive oxygen species were maximally generated between 0.5 to 1.5 hr, while damage to DNA expressed as 8-hydroyguanine adducts occurred between 4 and 6 hours. Both of these in vitro studies demonstrate that phosphine or its reaction products derived from AlP can generate DNA damage and that expression of these effects is delayed, probably indirectly and dependent on generation of hydrogen peroxides (Hsu et al., 1998). No completed long term animal studies were noted in this review regarding carcinogenicity. One preliminary report (U.S. EPA, 1998) noted no carcinogenic effects in rats chronically exposed to an inhaled dose of 3 ppm phosphine after one year. One human epidemiologic study of grain worker mortality (Alavanja et al., 1987a, b) shows an excess of cancers of the lymphatic and hematopoietic system in this occupational setting where exposure to phosphine and other chemicals and biologic agents occurs. One study of animal teratogenicity (Newton et al., 1993) with exposure concentrations as high as 4.9 ppm during days 6–15 of gestation in rats showed neither maternal toxicity nor developmental toxicity. No other reproductive endpoint studies were available for review.

104.10  Treatment of poisoning There is no current medical standard of treatment for acute phosphine intoxication. In general, support of vital functions, prevention and/or treatment of shock, and early gastric lavage for ingested poison are suggested (Singh et al., 1985). However, water based gastric lavage may enhance the reactivity of ingested aluminium phosphide tablets. Few clinical research efforts have been devoted to evaluation of antiperoxidants use (Chugh et al., 1996; Gupta and Ahlawat, 1995) such as magnesium sulfate for treatment of acute intoxication. More recently, the oral administration of the anti-ischemic drug trimetazidine, which works through a metabolic mechanism of decreasing the production of oxygen-derived free radicals and stimulating the oxidative metabolism of glucose (de Leiris and Boucher, 1993; Lopaschuk, 1998), was temporally associated with clinical improvement in a case report of occupational inhalation exposure to phosphine gas from aluminium phosphide (Dueñas et al., 1999). Shadnia et al. (2005) reported their recent experience with one phospine treatment case showed that rapid prevention of absorption by coconut oil might be helpful. Finally, there is clear need

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to fully evaluate use of antioxidants as potential therapeutic agents in light of the current toxicologic findings.

104.11  Regulatory notes (Exposure guidelines) NIOSH REL: TWA 0.3 ppm (0.4 mg/m3), STEL 1 ppm (1.4 mg/m3) OHSHA PEL: TWA 0.3 ppm (0.4 mg/m3) 1993-1994 AGGIH TLV: 0.3 ppm (0.42 mg/m3) TWA, 1 ppm (1.4 mg/m3) STEL Revised IDHL (immediately dangerous to health or life): 50 ppm (NIOSH, 1996) Acute reference dose (RfD) was established as 0.018 mg/kg/day (U.S. EPA, 1998). In this document, the risk assessment process identified concerns with respect to handlers (applicators) performing aeration and fumigation tasks without the use of respiratory protective equipment. Similar concerns were identified with respect to the potential risks of nonapplicators (in the vicinity of the application). A Memorandum of Agreement amending the aluminium phosphide RED was subsequently issued by the E.P.A., which outlined a series of provisions intended to mitigate risks (U.S. EPA, 2001). These provisions include fumigant exposure monitoring studies, improved training and education for applicators, the development of fumigation management plans, and plans to review and determine whether current exposure standards for phosphine should be lowered from the current limit of 0.3 ppm. The chronic reference dose was found to be 0.0113 mg/kg/day. Earlier EPA established RfD for phosphine was 0.0003 mg/kg/d based on body weight and clinical para­ meters (U.S. EPA, 1995).

104.12  Summary and comments Phosphine is a toxicant gas with strong reducing properties capable of chemical and biologic oxidant effects. The signature threshold for lethality over a narrow dose range and slow evolution of mortality at lower doses indicates that the chemical induces a cumulative biologic oxidant cascade involving progressive alteration of a number of critical biologic endpoints. The critical threshold for these effects may be moderated by environmental-chemical interactions affecting conditions of exposure. As O. R. Klimmer (1969) said “It is a most peculiar poison.”

References Abdner-Rahman, H. (1999). Effect of aluminum phosphide on blood glucose level. Vet. Hum. Toxicol. 41, 31–32. Abdner-Rahman, H. (2000). Aluminum phosphide fatalities, new local experience. Med. Sci. Law 40, 164–168.

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Alavanja, M. C., Malker, H., and Hayes, R. B. (1987a). Occupational cancer risk associated with the storage and bulk handling of agricultural foodstuff. J. Toxicol. Environ. Health 22, 247–254. Alavanja, M. C., Rush, G. A., Stewart, P., and Blair, A. (1987b). Proportionate mortality study of workers in the grain industry. J. Natl. Cancer I. 78, 247–252. Banjaj, R., and Wasir, H. S. (1988). Epidemic caluminium phosphide poisoning in northern India. Lancet 1, 820–821. Barbosa, A., and Bonin, A. M. (1994). Evaluation of phosphine genotoxicity at occupational levels of exposure in New South Wales, Australia. Occup. Environ. Med. 51, 700–705. Barbosa, A., Rosinova, E., Dempsey, J., and Bonin, A. M. (1994). Determination of genotoxic and other effects in mice following short term repeated-dose and subchronic inhalation exposure to phosphine. Environ. Mol. Mutagen. 24, 81–88. Brautbar, N., and Howard, J. (2002). Phosphine toxicity: report of two cases and review of the literature. Toxicol. Ind. Health 18, 71–75. Bogle, R. G., Theron, P., Brooks, P., Dargan, P. I., and Redhead, J. (2006). Aluminium phosphide poisoning. Emerg. Med. J. 23, e3 (http://www. emjonline.com/cgi/content/full/23/1/e3). doi: 10.1136/emj. 2004. 015941. Bolter, C. J., and Chefurka, W. (1990). Extramitochondrial release of hydrogen peroxide from insect and mouse liver mitochondria using the respiratory inhibitors phosphine, myxothiazol, and antimycin and spectral analysis of inhibited cytochromes. Arch. Biochem. Biophys. 278, 65–72. Bond, E. J., Robinson, J. R., and Buckland, C. T. (1969). The toxic action of phosphine: Absorption and symptoms of poisoning in insects. J. Stored Prod. Res. 5, 289–298. Buchanan, J. W., and Hanrahan, R. J. (1970). The radiation chemistry of phosphine–ammonia mixtures in the gas phase. Mutat. Res. 44, 206–304. Burgess, J. L., Morrissey, B., Keifer, M. C., and Robertson, W. O. (2000). Fumigant-related illnesses: Washington State’s five-year experience. Clin. Toxicol. 38, 7–14. Burgess, J. L. (2001). Phosphine exposure from a methamphetamine laboratory investigation. J. Toxicol. Clin. Toxicol. 39, 165–168. Cha’on, U., Valmas, N., Collins, P. J., Reilly, P. E. B., Hammock, B. D., and Ebert, P. R. (2007). Disruption of iron homeostasis increases phosphine toxicity in Caenorhabditis elegans. Toxicol. Sci. Offic. J. Soc. Toxicol. 96(1), 194–201. Chaudhry, M. Q. (1997). A review of the mechanisms involved in the action of phosphine as an insecticide and phosphine resistance in stored-product insects. Pestic. Sci. 49, 213–228. Chaudhry, M. Q., and Price, N. R. (1990). A spectral study of the biochemical reactions of phosphine with various haemproteins. Pestic. Biochem. Physiol. 36, 14–21. Chin, K. L., Meaklim, M. J., Scollary, G. R., and Leaver, D. D. (1992). The interaction of phosphine with haemoglobin and erythrocytes. Xenobiotica 22, 599–607. Chugh, S. N., Arora, V., Sharma, A., and Chugh, K. (1996). Free radical scavengers and lipid peroxidation in acute aluminum phosphide poisoning. Indian J. Med. Res. 104, 190–193. Chugh, S. N., Kolley, T., Kakkar, R., Chugh, K., and Sharma, A. (1997). A critical evaluation of anti-peroxidant effect of intravenous magnesium in acute aluminium phosphide poisoning. Magnes. Res. 10, 225–230. Chugh, S. N. (2003). Aluminium phosphide poisoning. In “API Textbook of Medicine” (S. N. Shah ed.), 7th ed, pp. 1272–1274. The Association of Physicians of India, Mumbai.

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Dévai, I., Felföldy, L., Wittner, I., and Plösz, S. (1998). Detection of phospine: New aspects of the phosphorous cycle in the hydrosphere. Nature 333, 343–345. Dueñas, A., Pérez-Castrillon, J. L., Cobos, M. A., and Herreros, V. (1999). Treatment of the cardiovascular manifestations of phosphine poisoning with trimetazidine, a new antiischemic drug. Am. J. Emerg. Med. 17, 219–220. Ellenhorn, M. J. (1999). “Medical Toxicology, Diagnosis and Treatment of Human Poisoning,” 2nd ed., Lippincott, Williams and Wilson, Baltimore, MD. Garry, V. F., Danzl, T. J., Nelson, R. L., Cervenka, J., Krueger, L. A., Griffith, J., and Whorton, E. (1989). Human genotoxicity: Pesticide applicators and phospine. Science 246, 251–255. Garry, V. F., Nelson, R. L., Danzl, T. J., Cervenka, J., Krueger, L. A., Griffith, J., and Whorton, E. (1990). Human genotoxicity in phosphineexposed fumigant applicators. Prog. Clin. Biol. Res. 340C, 367–376. Garry, V. F., Danzl, T. J., Tarone, R., and Griffith, J. (1992). Chromosome rearrangements in fumigant appliers: possible relationship to nonHodgkin’s lymphoma risk. Cane. Epi. Biomark. Prev. 1, 287–291. Garry, V. F., Good, P. E., Manivel, C., and Perl, D. (1993). Investigation of a fatality from nonoccupational aluminum phosphide exposure: Measurement of aluminum in tissue and body fluids as a marker of exposure. J. Lab. Clin. Med. 122, 739–747. Garry, V. F., Tarone, R. E., Long, L., Griffith, J., Kelly, J. T., and Burroughs, B. (1996). Pesticide appliers with mixed pesticide exposure: G-banded analysis and possible relatonship to non-Hodgkin’s lymphoma. Cane. Epi. Biomark. Prev. 5, 11–16. Glindemann, D., Stottmeister, U., and Bergmann, A. (1996). Free phosphine from the anaerobic biosphere. Environ. Sci. Pollut Res. Intern. 3, 17–19. Green, A. R., Sheldon, S., and Banks, H. J. (1984). The flammability limit of pure phosphine-air mixtures at atmospheric pressure. In Controlled Atmosphere and Fumigation in Grain Storage (B. E. Ripp ed.) Vol. 5, pp. 433–451. Elsevier, Amsterdam. Gupta, S., and Ahlawat, S. K. (1995). Aluminum phosphide poisoning: A review. J. Toxicol. Clin. Toxicol. 33, 19–24. Heyndrickx, A., Van Peteghem, C., Van Den Heede, M., and Lauwaert, R. (1976). A double fatality with children due to fumigated wheat. Eur. J. Toxicol. 9, 113–118. Hsu, C.-H., Quistad, G. B., and Casida, J. E. (1998). Phosphine induced oxidative stress in Hepa lclc7 cells. Toxicol. Sci. 46, 204–210. Hsu, C.-H., Han, B.-C., Liu, M.-Y., Yeh, C.-Y., and Casida, J. E. (2000a). Phosphine-induced oxidative damage in rats: attenuation by melatonin. Free Radic. Biol. Med. 28, 636–642. Hsu, C.-H., Chi, B.-C., and Casida, J. E. (2002b). Melatonin reduces phosphine-induced lipid and DNA oxidation in vitro and in vivo in rat brain. J. Pineal Res. 32, 53–58. Hsu, C.-H., Chi, B.-C., Liu, M.-Y., Li, J.-H., Chen, C.-J., and Chen, R. Y. (2002). Phosphine-induced oxidative damage in rats: role of gluta­ thione. Toxicology 179, 1–8. Kashi, K. P., and Chefurka, W. (1976). The effect of phosphine on the absorption and circular dichroic spectra of cytochrome c and cytochrome oxidase. Pestic. Biochem. Physiol. 6, 350–362. Kaushik, R. M., Kaushik, R., and Mahajan, S. K. (2007). Subendocardial infarction in a young survivor of aluminium phosphide poisoning. Hum. Exp. Toxicol. 26(5), 457–460. Kligerman, A. D., Bryant, M. F., Doerr, C. L., Erexson, G. L., Kwanyuen, P., and McGee, J. K. (1994a). Cytogenic effects of phosphine inhalation by rodents: I. Acute 6 hour exposure of mice. Environ. Mol. Mutagen. 23, 186–189.

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Kligerman, A. D., Bishop, J. B., Erexson, G. L., Price, H. C., O’Connor, R. W., Morgan, D. L., and Zeiger, E. (1994b). Cytogenic and germ cell effects of phosphine inhalation by rodents. II. Subacute exposures to rats and mice. Environ. Mol. Mutagen. 24, 301–306. Klimmer, O. R. (1969). Beitrag zur Wirkung des Phosphorwasserstoffes (PH3). Zur Frage der sog chronischen Phosphorwasserstoffvergiftung. Arch. Toxikol. 24, 164–187 [in German]. Klimmer, O. R. (1970). Akute Vergiftungen durch Insektizide und Herbizide. Z. Allgemeinmedizin 46, 1731–1734 [in German]. Kurzbauer, H., and Kiesler, A. (1987). Occupational phosphine poisoning. Neurologia i Neurochirurgia Polska 20, 415–417. Lam, W. W., Toia, R. F., and Casida, J. E. (1991). Oxidatively initiated phosphorylation reactions of phosphine. J. Agric. Food Chem. 39, 2274–2278. Lopaschuk, G. D. (1998). Treating ischemic heart disease by pharmacologically improving cardiac energy metabolism. Am. J. Cardiol. 82, 14K–17K. Meister, R. T. (1999). “Farm Chemicals Handbook.” Meister, Wikkoughby. Mehrpour, O., Alfred, S., Shadnia, S., Keyler, D. E., Soltaninejad, K., Chalaki, N., and Sedaghat, M. (2008). Hyperglycemia in acute aluminum phosphide poisoning as a potential prognostic factor. Hum. Exp. Toxicol. 27(7), 591–595. Misra, U. K., Bhargave, S. K., Nag, D., Kidwai, M. M., and Lai, M. M. (1988a). Occupational phosphine exposure in Indian workers. Toxicol. Lett. 42, 257–263. Misra, U. K., Tripafhi, A. K., Pandey, R., and Bhargwa, B. (1988b). Acute phosphine poisoning following ingestion of aluminum phosphide. Hum. Toxicol. 7, 343–345. Nakakita, H. (1987). The mode of action of phosphine. J. Pestic. Sci. 12, 299–309. Newton, P. E., Schroeder, R. E., Sullivan, J. B., Busey, W. M., and Banas, D. A. (1993). Inhalation toxicity of phosphine in the rat: Acute, subchronic, and developmental. Inhal. Toxicol. 5, 223–239. NIOSH Pocket Guide to Chemical Hazards. (1996). “Documentations for Immediately Dangerous to Life or Health Concentrations (IDLH): Phosphine.” NIOSH Alert. (1999). Preventing Phosphine Poisoning and Explosions During Fumigation. National Occupational Hazard Survey Publication, September 1999. Olson, K. R. (2006). Phoshine and phosfides. In “Poisoning & Drug Overdose: By the Faculty, Staff and Associates of the California Poison Control System” (By K. R. Olson, I. B. Anderson, California Poison Control System; Contributor K. R. Olson, I. B. Anderson, N. L. Benowitz), 5th ed., pp. 307-308. McGraw-Hill Professional. Omae, K., Ishizuka, C., Nakashima, H., Sakurai, H., Yamazaki, K., Mori, K., Shibata, T., Kanoh, H., Kudo, M., and Tati, M. (1996). Acute and subacute inhalation toxicity of highly purified phosphine (PH3) in male ICR mice. J. Occup. Health 38, 36–42. Osadchenko, I. M., and Tomilov, A. P. (1969). Phosphorous hydrides. Russian Chem. Rev. 33, 495–504. Pazynich, V. M., Mazur, I. A., Podlozny, A. V., Chinchevich, V. I., and Mandrichenko, B. E. (1984). Experimental substantiation and prediction of time related maximum permissible concentration of phosphine in the air. Gig. Sanit. 1, 13–15 [in Russian]. Potter, W. T., Rong, S., Griffith, J., White, J., and Garry, V. F. (1991). Phosphine mediated Heinz Body formation and hemoglobin oxidation in human erythrocytes. Toxicol. Lett. 57, 37–45. Potter, W. T., Garry, V. F., Kelly, J. T., Taronef, R., Griffith, J., and Nelson, R. L. (1993). Radiometric assay of red cell and plasma cholinesterase

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in pesticide appliers from Minnesota. Toxicol. Appl. Pharmacol. 119, 150–155. Price, N. R. (1980). Some aspects of the inhibition of cytochrome c oxidase by phosphine in susceptible and resistant strains of Rhyzopertha Dominicia. Insect Biochem. 10, 147–150. Robinson, J. R., and Bond, E. J. (1970). The toxic action of phosphine: Studies with 32PH3; terminal residues in biological materials. J. Stored Prod. Res. 6, 133–146. Shadnia, S., Rahimi, M., Pajoumand, A., Rasouli, M. H., and Abdollahi, M. (2005). Successful treatment of acute aluminium phosphide poisoning: possible benefit of coconut oil. Hum. Exp. Toxicol. 24(4), 215–218. Shadnia, S., Mehrpour, O., and Abdollahi, M. (2008). Unintentional poisoning by phosphine released from aluminum phosphide. Hum. Exp. Toxicol. 27(1), 87–89. Singh, S., Dilawari, J. B., Vashist, R., Malhotra, H. S., and Sharma, B. K. (1985). Aluminum phosphide ingestion. Br. Med. J. (Clin. Res.) 290, 1110–1111. Singh, S., Singh, D., Wig, N., Jit, I., and Sharma, B. K. (1996). Aluminum phosphide ingestion—A clinico-pathologic study. J. Toxicol. Clin. Toxicol. 34, 703–706. Siwach, S. B., Dua, A., Sharma, R., Sharma, D., and Mehla, R. K. (1995). Tissue magnesium content and histopathological changes in nonsurvivors of aluminium phosphide poisoning. J. Assoc. Phys. Ind. 43, 676–678. Sudakin, D. L. (2005). Occupational exposure to aluminium phosphide and phosphine gas? A suspected case report and review of the literature. Hum. Exp. Toxicol. 24, 27–33. Trimborn, H., and Klimmer, O. R. (1962). Experimentelle untersuchungen über chemische veränderungen des blutfarbstoffs in vitro durch phosphorwasser-stoff. Arch. Int. Pharmacodyn. CSSSVII, 331–347.

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U.S. Department of Health and Human Services. (1993). “Hazardous Substances Data Bank” (HSDB, online database). National Toxicilogy Information Program, National Library of medicine, Bethesda. U.S. EPA. (1995). “Integrated Risk Information System (IRIS) on Phosphine.” Environmental Criteria and Assessment Office, Office of Health and Environmental Assessment, Office of Research and Development, Cincinnati, OH. U.S. EPA. (1998). “Prevention, Pesticides And Toxic Substances (7508C). Reregistration Eligibility Decision (RED) Aluminum and Magnesium Phosphide.” EPA 738-R-98-017. Van Wazer, J. R. (1958). “Phosphorus and Its Compounds, Vol. I. Chemistry,” Interscience Publishers, New York. Waritz, R. S., and Brown, R. M. (1975). Acute and subacute inhalation toxicities of phosphine, phenylphosphine and triphenylphosphine. Am. lnd. Hyg. Assoc. J. 36, 452–458. Willers-Russo, L. J. (1999). Three fatalities involving phosphine gas, produced as a result of methamphetamine manufacturing. J. Forensic Sci. 44, 647–652. Wilson, R., Lovejoy, F. R., Jaeger, R. J., and Landrigan, P. L. (1980). Acute phosphine poisoning aboard a grain freighter. JAMA, J. Am. Med. Assoc. 244, 148–150. Woller, C. R. (1965). Aliphatic Compounds of Some Elements. In “Chemistry of Organic Compounds” (W. B. Saunders, ed.), pp. 317–323. World Health Organization (1988). Phosphine and selected metal phosphides. Environ. Health Criteria 73, 17–19. Zuryn, S., Kuang, J., and Ebert, P. (2008). Mitochondrial modulation of phosphine toxicity and resistance in Caenorhabditis elegans. Toxicol. Sci. Offic. J. Soc. Toxicol. (Toxicol. Sci.) 102(1), 179–186.

Chapter 105

Methyl Bromide Vincent J. Piccirillo and Amanda L. Piccirillo VJP Consulting, Inc., Ashburn, Virginia

105.1  Introduction Methyl bromide is a broad spectrum pesticide primarily used for soil fumigation, commodity/quarantine treatment, and structural fumigation. Since human exposure is more likely to occur by the inhalation route, the majority of toxicologic evaluations for methyl bromide are inhalation studies. Ingestion of fumigated commodities is a secondary route of human exposure. Chronic dietary studies in rats and dogs show no concern for long-term oral ingestion of methyl bromide. This chapter briefly describes many of the published studies and elaborates the results from a number of contemporaneous unpublished methyl bromide toxi­ city studies by the oral and inhalation routes, which were conducted to support pesticide registration and other regulatory needs of the U.S. Environmental Protection Agency as well as state and international regulatory bodies. A primary focus in this chapter is methyl bromide-induced neurotoxicity. From a risk characterization standpoint, clinical observations of neurotoxicity are considered the primary endpoint of concern from inhalation exposure. Review of the overall toxicity shows that methyl bromide-induced toxicity is a function of both the concentration and the duration of exposure. This is an important consideration for human exposure assessments.

105.2  Chemical properties and pesticidal uses of methyl bromide Methyl bromide (CH3Br, bromomethane, CAS no. 74-83-9) is a colorless, odorless gas at normal temperature and pressure, and is produced by the interaction of methanol Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

(CH3OH) and hydrogen bromide (HBr). Methyl bromide is made commercially but also is produced naturally by marine algae and other plants and as a by-product of the combustion of plant materials (i.e., forest fires). Under increased pressure or below 3°C, methyl bromide is a clear to straw-colored liquid and is usually shipped as a liquified, compressed gas. Methyl bromide has a boiling point of 38.5°F and is nonflammable in air. Methyl bromide formulations contain chloropicrin, an irritant and lacrimator, as a warning agent. Methyl bromide is a broad-spectrum pesticide primarily used for soil fumigation, commodity/quarantine treatment, and structural fumigation. It is also used as an intermediate in the manufacture of other chemicals. Methyl bromide has been used as a fumigant for more than 50 years and is strictly controlled by the U.S. Environmental Protection Agency (U.S. EPA) under the Federal Insecticide, Fungicide, and Rodenticide Act. Its application and use are also controlled by various state regulatory authorities. For soil fumigation, methyl bromide is injected directly into the soil, which is then covered with a variety of tarping materials. The tarps are sealed, kept in place for several days, and then removed. Soil fumigation with methyl bromide enhances the quality of the crops and increases yield by eliminating fungal diseases, nematodes, weed seeds, and other soil-borne pests. The primary crops grown in methyl bromide-treated soil are peppers, strawberries, tomatoes, and grapes. Methyl bromide is widely used for fumigating postharvest commodities, including wheat and cereals, spices, nuts, and dried or fresh fruits, to eradicate pest infestations. Fumigation typically occurs where the commodities are stored, such as in ship holds, grain elevators, warehouses, 2267

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special fumigation chambers, and on shipping piers/docks. Commodity fumigation typically involves the use of specially designed and permanently installed chambers into which the methyl bromide is released. After treatment, mechanical ventilation continuously aerates the commodity until concentration of methyl bromide in the vented air is at established safety levels. Another type of commodity fumigation involves sealing the commodity under a tar­ paulin followed by injection of methyl bromide. Aeration and ventilation occurs after the tarpaulin is removed. In structural fumigation, all openings in the structure are sealed. All types of commercial and residential structures may be fumigated with methyl bromide to control or eradicate pests, such as termites. The structure is covered by a “tent” or tarpaulin, and methyl bromide gas is released inside the structure. After a specified period, the tarpaulin is removed and the structure is aerated until concentration of methyl bromide inside the structure reaches safe levels. Methyl bromide may only be applied and used by professional, certified applicators. All persons working with methyl bromide are required to be knowledgeable about its hazards and trained in the use of required respiratory protection equipment, detector devices, and emergency proced­ ures. Applicators and other persons in the fumigation area must wear appropriate personal protective equipment (PPE) as required by the label and U.S. EPA regulations. Such PPE typically includes full eye/face shields, safety shoes, and respirators. If the concentration of methyl bromide in the work area exceeds established safety levels, all persons in the fumigated area must wear an approved, selfcontained breathing apparatus or evacuate the area. The placarding and posting of warning notices at all entrances to an area undergoing fumigation is required. No one is permitted in a structure or area undergoing fumigation, unless involved in the fumigation and wearing appropriate PPE. Re-entry into fumigated areas or structures is ­prohibited until the air concentration of methyl bromide is shown to be at safe levels. Individuals living in close proximity to fumigated fields, greenhouses, or structures are unlikely to be exposed to unsafe levels of methyl bromide because of the application restrictions and the rapid dissipation of methyl bromide in the atmosphere. Additional regulatory controls further limit this possibility. For ­example, in California, maximum air concentration levels have been established for state-mandated buffer zones ­surrounding fumigated areas.

105.3  Toxicology of methyl bromide The toxicology of methyl bromide has been extensively reviewed (ATSDR, 1991; WHO, 1995). This chapter briefly describes many of the published studies and elaborates the results from a number of contemporaneous

unpublished toxicity studies with methyl bromide, which were conducted to support the EPA reregistration and to provide specific data to meet various state registration/ regulatory requirements. A primary focus in this chapter is methyl bromide-induced neurotoxicity. From a risk characterization standpoint, clinical observations of neurotoxicity are considered as the primary endpoint of concern from inhalation exposure. Reviewing the overall toxicity of methyl bromide shows that methyl bromide-induced toxi­city is a function of both the concentration and the dur­ation of exposure.

105.3.1  Acute Toxicity 105.3.1.1  Oral The acute LD50 for methyl bromide in rats was reported as 214 mg/kg (Danse et al., 1984). Prior to conducting longer-term toxicity studies, an acute oral toxicity study was conducted that compared liquid methyl bromide to a microencapsulated form. Similar oral toxicity was noted for both forms of methyl bromide; the oral LD50 values were 104 mg/kg for liquid methyl bromide and 133 mg/kg for microencapsulated methyl bromide (Kiplinger, 1994). In beagle dogs, 500 mg/kg produced severe signs of toxicity and vomiting followed by death within 24 h of dosing. A 50-mg/kg dose elicited signs of toxicity and vomiting of reddish material but no deaths. At low doses of 3 and 5 mg/kg, dogs vomited shortly after dosing. An oral LD50 study could not be conducted since dogs vomited the dose (Naas, 1990).

105.3.1.2  Inhalation Overt toxicity (i.e., death) from acute inhalation exposures to methyl bromide has been extensively evaluated in rodents (Alexeeff and Kilgore, 1985; Irish et al., 1940; Japanese Ministry of Labour, 1992; Zwart, 1988). The majority of acute inhalation studies in mice and rats demonstrate that methyl bromide exposure-related effects and mortality are a function of both the concentration and the duration of exposure. Inhalation LC50 values for methyl bromide in mice have been reported as 1700 ppm for a 30-min expos­ ure (Bakhishev, 1973), 1200 ppm for a 60-min exposure (Alexeeff and Kilgore, 1985), 397 ppm for a 120-min exposure (Balander and Polyak, 1962), and 405 ppm for a 240-min exposure (Yamano, 1991). Similarly, inha­lation LC50 values for rats were 2833 ppm for 30-min expos­ ure (Bakhishev, 1973), 1880 ppm for a 60-min exposure (Zwart, 1988; Zwart et al., 1992), 781 ppm for a 240-min exposure (Kato et al., 1986), and 302 ppm for an 8-h expos­ure (Honma et al., 1985). A number of acute inhalation studies provide results that clearly demonstrate the concentration and the duration of exposure relationship for methyl bromide.

Chapter | 105  Methyl Bromide

In an acute inhalation study, groups of F344 rats were exposed to methyl bromide at concentrations of 150, 225, 338, 506, 760, or 1140 ppm for 4 h (Japanese Ministry of Labour, 1992). This single exposure resulted in decreased locomotor activity, ataxia, nasal discharge, lacrimation, diarrhea, irregular breathing, and bradypnea in rats exposed to 338 ppm and greater. No clinical signs of toxicity were evident in animals exposed to 225 ppm methyl bromide or less. Histologic evaluations revealed metaplasia of the olfactory epithelium for rats exposed to 225, 338, and 506 ppm methyl bromide. Honma et al. (1985) conducted a series of acute inhal­ ation toxicity studies to evaluate methyl bromide-induced effects on locomotor activity, body temperature, body weight gain, and enhancement of thiopental-induced sleep. Rats were exposed to methyl bromide concentrations of 63, 125, 188, or 250 ppm for 8 h. Enhanced thiopental sleep potentiation, measured by time to loss of righting reflex upon thiopental injection, was noted at 63 ppm and greater. Body weight gain and body temperature were decreased in rats exposed to concentrations of 125 ppm and greater. Neurotoxicity, indicated by reduced locomotor activity, was seen at concentrations of 188 and 250 ppm. These effects were reversible within 24 h of exposure. In a study evaluating histologic changes from acute inhalation of methyl bromide (Hurtt et al., 1987), Fischer 344 rats were exposed to methyl bromide concentrations of 0, 90, 175, 250, or 325 ppm on a 6-h/day, 5 consecutive day regimen. Diarrhea was noted by the end of the 2nd day of exposure for 250- and 325-ppm animals. By the end of the third exposure, animals from these groups showed ataxia. Two of the 325-ppm rats exhibited tremors and/or convulsions during the fourth exposure. Subsequently, three animals from this group succumbed after the fourth exposure. Clinical signs of neurotoxicity were not observed in the 250- and 325-ppm animals after a single (or second) expos­ ure to methyl bromide. Irish et al. (1940) exposed rats and rabbits to methyl bromide concentrations ranging from 108 to 12,850 ppm. Study results showed clear concentration and expos­ ure duration dependence. Rabbits tolerated exposure to 220 ppm methyl bromide for 20 h but exposure at this concentration for 32 h resulted in 100% mortality. At 2570 ppm, rabbits survived exposure at 1 h but 100% mortality was observed after 2.2 h of exposure. Groups of mice were exposed for 4 h via whole body exposure to methyl bromide atmospheric concentrations of 100, 150, 225, 338, 506, or 760 ppm (Japanese Ministry of Labour, 1992). Concentration-dependent clinical signs of toxicity consisting of decreased locomotor activity, ­tremors, convulsions, diarrhea, dyspnea, and bradypnea were seen at concentrations of 506 and 760 ppm. The acute inhalation no-obvious-adverse-effect level (NOAEL) for clinical signs of neurotoxicity in mice exposed to methyl bromide was 338 ppm.

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Alexeeff and Kilgore (1985) conducted a series of 1-h inhalation exposure studies in which mice were exposed to methyl bromide concentrations ranging from 225 to 1530 ppm. Based on all signs of neurotoxicity, the NOEL was 560 ppm. In a series of inhalation studies (Newton, 1994b), beagle dogs received 1 to 4 days exposure to methyl bromide. The purpose of this study was to determine tolerable inhalation exposure levels to be used in a 4-week inhalation study. In the initial phase of the study, three males and three females were exposed for 6–7 h to methyl bromide concentrations of 233 (one male), 314 (one male, one female), 345/350 (one male, one female), or 394 ppm (one female). Signs of toxicity were observed at all concentrations, therefore, the 1-day NOAEL was 233 ppm. In the second phase of the study, dogs were exposed to either 55 ppm (one male, one female), 156 (one male, one female), 268 ppm (one male, two females), or 283 ppm (two males, one female) for 7 h/day for up to 4 days. The 268-ppm and 283-ppm dogs were exposed to methyl bromide for 2 days and developed clinical signs of toxicity. Therefore, the 2-day NOAEL for beagle dogs exposed to methyl bromide was 268 ppm. The 55-ppm and 156-ppm dogs were exposed 7 h/day for 4 consecutive days. No effects were seen in either the 55- or 156-ppm dogs during days 1 and 2 of exposure. However, the 156-ppm animals showed decreased activity during exposure on days 3 and 4, and irregular gait during the postexposure period on day 4.

105.3.2  Subchronic Toxicity 105.3.2.1  Oral In a subchronic toxicity study (Danse et al., 1984), Wistar rats were dosed with methyl bromide via gavage at doses of 0, 0.4, 2, 10, or 50 mg/kg/day for 90 days. At 50 mg/kg/day, marked, diffuse hyperplasia of the epithelium of the forestomach was seen in all animals. Squamous cell carcinoma of the forestomach was diagnosed in 13 of 20 animals receiving methyl bromide at 50 mg/kg/day. Upon subsequent evaluation of the histology slides, it was concluded that the forestomach lesions represented inflammation and hyperplasia rather than malignant lesions. Inflammatory lesions of the forestomach were also seen in animals treated with 2 and 10 mg/kg/day.

105.3.2.2  Inhalation Table 105.1 summarizes the results from a number of subchronic inhalation studies in rats and mice. For comparison purposes, study results have been presented to show the overall study NOEL, the NOEL for neurotoxicity, and the LOEL for neurotoxicity. Beagle dogs were exposed to either 55 ppm (one male, one female), 156 (one male, one female), 268 ppm (one

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Table 105.1  Summary of Subchronic Inhalation Toxicity Studies with Methyl Bromide Study

Species (strain)

Exposure (h/days/weeks)

Overall NOEL

Neurotoxicity NOEL (effect)

Neurotoxicity LOEL

Subchronic neurotoxicity (Norris et al., 1993)

Rat (SD)

6/5/13

30 ppm

30 ppm

70 ppm

Subchronic toxicity (NTP, 1992)

Rat (SPF Wistar)

6/5/3; 6/7/1

18 ppm

18 ppm

51 ppm

Subchronic toxicity (Kato et al., 1986)

Rat (SD)

4/5/6

150 ppm

200 ppm (dec. body weight)

300 ppm

Subchronic toxicity (Haber et al., 1985) (NTP, 1992)

Rat (F344/N)

6/5/13

30 ppm

60 ppm (dec. body weight)

120 ppm

Subchronic toxicity (Japanese Ministry of Labour, 1992)

Rat (F344/DuCrj)

6/5/13

7.5 ppm

117 ppm (clinical pathology)

293 ppm

Subchronic toxicity (Wilmer et al., 1983)

Rat (Wistar)

6/5/13

6.4 ppm

42 ppm (liver pathology)

No neurotoxicity

Subchronic toxicity (NTP, 1992)

Mouse (B6C3F1)

6/5/13

20 ppm

80 ppm (hematology) 120 ppm

Subchronic toxicity (Japanese Ministry, 1992)

Mouse (Crj:BDF1)

6/5/13

30 ppm

60 ppm (body weight, hematology, urinalysis)

male, two females), or 283 ppm (two males, one female) for 7 h/day for up to 4 days (Newton, 1994b). Clinical signs of toxicity were seen in the 268- and 283-ppm dogs after two exposures. The 156-ppm animals showed decreased activity during exposure on days 3 and 4, and irregular gait during the postexposure period on day 4. The 55-ppm concentration was the 4-day NOEL. In a 4-week inhalation study (Newton, 1994a), beagle dogs (four/sex/group) were exposed for 5 days/week, 7 h/day to methyl bromide at concentrations of 0, 5, 10, 25, 50, or 100 ppm. No clinical evidence of neurotoxicity was seen in any group throughout the 4 weeks of exposure. After 4 weeks, four of the controls (two females and two males) and all dogs in the 5-ppm group continued the test for an additional 2 weeks and the exposure concentration for the 10-ppm group was increased to 150 ppm. Dogs exposed to 150 ppm methyl bromide showed severe body weight loss over the first few days of exposure. After five or six exposures, evaluation of the 150-ppm animals by a veterin­ ary neurologist revealed ataxia, a base-wide stance, intention tremor, nystagmus, marked depression, and inability (unwillingness) to stand and perform postural responses. Due to the severity of these effects, the dogs were sacrificed. Neurologic evaluation revealed no treatment-related neurologic effects for dogs exposed at the lower methyl bromide concentrations. Microscopic findings were limited to the 150-ppm group in which significant neurologic

No neurotoxicity

effects were seen and consisted of vacuoles in the granular layer of the cerebellum. In a 6-week inhalation study (Schaefer et al., 2002), four male and four female beagle dogs were exposed to methyl bromide by whole body inhalation at concentrations of 0, 5.3, 10, and 20 ppm for 5 days/week, 7 h/day. There were no methyl bromide-related effects on mortality, clinical signs, body weights, food consumption, body temperature, functional observational battery, locomotor activity, or macroscopic and microscopic pathology. One male and one female at the 20-ppm dose and one male at the 10-ppm dose demonstrated an absence of proprioceptive placing, although no evidence of weakness in motor strength or other signs of neurotoxicity were found. Therefore, the LOAEL was 10 ppm for male dogs and 20 ppm for female dogs. The NOAELs were 5.3 ppm for male dogs and 10 ppm for female dogs.

105.3.3  Genetic Toxicity Methyl bromide has been tested in numerous in vitro and in vivo genetic toxicity studies with variable results. Methyl bromide has been reported to induce mutagenic effects in bacterial tests with Salmonella typhimurium (TA100 and TA1535), Escherichia coli, and Klebsiella pneumoniae (Djalali-Behzad et al., 1981; Kramers et al.,

Chapter | 105  Methyl Bromide

1985b; Moriya et al., 1983; Simmon et al., 1977). No evidence of mutagenicity was seen when methyl bromide was tested in a modified Ames test using an in situ impingement test system, but a significant response was seen with the SOS repair test (Ong et al., 1987). A sex-related recessive lethal assay was conducted with male strain Oregon K Drosophila melanogaster (McGregor, 1981). The Drosophila were exposed to methyl bromide concentrations of 20 or 70 ppm for 5 h and then allowed to mate on days 1, 3, or 8 following exposure. The F1 progeny were mated brother to sister, 1 to 4 days after emergence. The resulting F2 generation was then examined for the absence of wild-type males. No compound-related increases in the frequency of lethal mutations in the F2 generation were noted. In a second sex-related recessive lethal assay, Drosophila melanogaster of the Berlin K strain were exposed to methyl bromide concentrations ranging from 18 to 192 ppm for varying exposure intervals (Kramers et al., 1985a,b). As was noted for the acute inhalation studies in mammalian species, mutagenic responses were related to both exposure concentration and duration. No increase in mutation frequency was seen in Drosophila exposed to 192 ppm for 6 h. Exposure at 155 ppm resulted in all flies dying during the 4th day of exposure. At lower concentrations, prolonged exposure resulted in mutagenic responses. Exposure at 125 ppm for 5 days (6 h per day) and at 50 ppm for 15 days (6 h per day) were considered mutagenic. Exposure of L5178Y mouse lymphoma cells to methyl bromide concentrations ranging from 7.7 to 7710 ppm resulted in dose-related increases in 6-thioguanine- and bromodeoxyuridine-resistant mutants (Kramers et al., 1985b). Sister chromatid exchanges (SCEs) and chromosomal aberrations in human lymphocytes exposed to methyl bromide were evaluated. Exposure of human lymphocyte cultures to an atmosphere of 4.3% methyl bromide for 100 s increased the frequency of SCEs from 10.0 to 16.8 per cell (Tucker et al., 1986). When human lymphocytes were treated with methyl bromide (0–24 g/ml) for 30 min, dose-related increases in SCEs and chromosomal aberrations were found. Metabolic activation (S9) significantly induced chromosomal aberrations (Garry et al., 1990). In a mitotic recombination assay in somatic cells (somatic wing spot assay), Drosophila melanogaster (third instar larvae trans-dihybrid for two recessive wing hair mutations) were exposed to methyl bromide vapor concentrations ranging up to 5140 ppm for 1 h. Wings of surviving adults were evaluated for the presence of cellular clones with malformed wing hairs. Methyl bromide-induced mitotic recombination as exhibited by the observation of small and large single as well as twin spots were observed (Katz, 1985, 1987). A rodent micronucleus (MN) study was conducted in BDF1 mice and F344 rats (10/sex/group) exposed via

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vapor inhalation to methyl bromide concentrations of 0, 154, 200, 260, 338, or 440 ppm for 6 h/day, 5 days/week for 2 weeks (Araki et al., 1995). Bone marrow of rats and peripheral blood of mice were evaluated for MN induction. In mice, significantly increased incidences of micronuclei in bone marrow polychromatic erythrocytes (PCE) were observed in males at 154 and 200 ppm and in females at 154 ppm; smaller increases in MN frequency were observed in normochromatic erythrocytes. Peripheral blood showed significant increases in MN at 200 ppm in males and 154 ppm in females. Due to excessive mortality, mice exposed to methyl bromide concentrations of 260 ppm and greater were not assayed. In rats, a statistically significant increase of MN in PCE was seen for males exposed at 338 ppm. A statistically insignificant increase of MN in PCEs was seen in female rats exposed at 260 and 338 ppm. Rats exposed at 400 ppm were not assayed due to excessive mortality. Methyl bromide was selected by the National Institutes of Occupational Safety and Health for evaluation in a Tier II Mutagenic Screening (McGregor, 1981). The testing program included: (1) unscheduled DNA synthesis (UDS) assay in human diploid fibroblasts; (2) sex-linked recessive lethal test in Drosophila melanogaster; (3) cytogenetic test in bone marrow cells of male and female rats; (4) sperm abnormality test in male mice; and (5) dominant lethal test in male rats. Summary results from this screening battery are as follows: Human diploid fibroblasts exposed for 3 h to methyl bromide concentrations of up to 70% in air over a minimal volume of culture medium in a UDS resulted in no increase in UDS. l Drosophila melanogaster exposed to methyl bromide concentrations of 20 or 70 ppm for 5 h did not have an increased frequency of sex-linked recessive mutations. l Cytogenetic analysis of bone marrow cells derived from male and female Sprague Dawley rats exposed for 1 or 5 days (7 hs/day) to methyl bromide concentrations of 20 or 70 ppm showed no treatment-related increases in the frequency of chromosomal aberrations in any of the methyl bromide-exposed groups. l Evaluation and characterization of sperm from male B6C3F1 mice exposed to 20 or 70 ppm methyl bromide for 7 h/day on five consecutive days then sacrificed 5 weeks later showed no significant increase in the frequency of abnormal sperm. l In the dominant lethal study, male Sprague Dawley rats were exposed to methyl bromide concentrations of 0 (air), 20, or 70 ppm for 7 h/day for 5 consecutive days, then allowed to breed with two virgin females weekly for 10 weeks. The females were sacrificed on day 14 after presumed mating. Examination of ovaries and uterine contents showed no evidence of genotoxicity. l

In a separate study, methyl bromide was evaluated for its ability to induce single-strand breaks in rat testicular DNA

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using alkaline elution techniques (Bentley, 1994). Groups of 10 male Fischer 344 rats were exposed to methyl bromide vapor concentrations of 0, 75, 150, or 250 ppm for 6 h/day over 5 consecutive days. The negative control group was exposed to room air only. A positive control group received a single intraperitoneal injection of 50 mg/kg methyl me­thanesulfonate in phosphate buffered saline. Five animals from each group were sacrificed at 1 and 24 h postexpos­ ure. Significant toxicity was seen in the 250 ppm rats. Two males from this group died and a third male was sacrificed in extremis with the 1-h post-treatment animals. Surviving rats showed decreased body weight and clinical signs of toxicity characterized by ataxia, spasms, diarrhea, lethargy, and prostration. At 150 ppm, the male rats showed slight body weight loss, nasal and/or ocular discharge, and wet and/or stained perineum during the 5-day exposure. A statistically significant increase in the mean elution rate of testicular cell DNA was observed at both sacrifice times only in rats exposed to the highly toxic 250-ppm concentration.

105.3.4  Developmental and Reproductive Toxicity In a developmental toxicity study (Sikov et al., 1981), pregnant female Wistar rats were exposed to methyl bromide concentrations of 0, 20, or 70 ppm for 7 h/day on gestation days 1–19. In addition, some groups were exposed pregestationally to 20 or 70 ppm for 3 weeks (5 days/week) immediately prior to mating. The distribution of dose groups (pregestational exposure concentration/gestational exposure concentration) was 0/0, 0/20, 0/70, 20/0, 20/20, 70/0, and 70/70 ppm. Cesarean sacrifice was performed on gestation day 19. No clinical evidence of maternal toxicity, fetotoxicity, or developmental toxicity was observed in any exposure scenario. Artificially inseminated New Zealand White rabbits (24 per group) were exposed daily on gestation days 1–24 to methyl bromide concentrations of 0 and 20 ppm; a group of inseminated rabbits exposed to 70 ppm methyl bromide were terminated due to excessive mortality and neurotoxicity characterized by convulsions and paresis in the hindlimbs seen after 1 week of treatment (Sikov et al., 1981). Control and 20-ppm-exposed rabbits were sacrificed on gestation day 30. No fetotoxicity or developmental toxicity was noted for the 20-ppm group. In probe studies (Breslin et al., 1990a), pregnant rabbits were exposed to methyl bromide concentrations of 0, 10, 30, or 50 ppm in one study, and concentrations of 0, 50, 70, or 140 ppm in a second study. Rabbits were exposed for 6 h/day on gestation days 7–19. Evidence of toxicity was observed only in the 140-ppm group does. All does exposed to methyl bromide at this concentration showed lethargy and decreased food consumption after eight

Hayes’ Handbook of Pesticide Toxicology

e­ xposures. Continued exposure signs of neurotoxicity were apparent and resulted in sacrifice of the does on gestation day 17. No apparent embryotoxicity was observed at any exposure level. The subsequent developmental toxi­ city study in rabbits was conducted in two phases (Breslin et al., 1990a,b). In the initial phase, pregnant New Zealand White rabbits were exposed for 6 h/day to methyl bromide concentrations of 0, 20, 40, and 80 ppm on days 7–19 of gestation. In the second phase, pregnant does were exposed to 0 or 80 ppm only. Caesarean delivery was performed on day 28 of gestation. In the first phase, maternal toxi­ city, evidenced by decreased body weight gain and clinical signs of neurotoxicity, was observed in three does from the 80-ppm group. Clinical signs consisted of right-sided head tilt, ataxia, slight lateral recumbency, and lethargy. In the second study, a significant decrease in body weight during gestation was the only evidence of maternal toxicity in the 80-ppm group. Developmental effects were only seen at 80-ppm group, a concentration showing significant maternal toxicity. In phase 1, fetal findings consisted of low incidences of omphalocele, hemorrhaging with or without hydrops (edema), retroesophogeal right subclavian artery, gall bladder agenesis, and fused sternebra. Fetal effects in phase 2 were limited to decreased fetal weight, hemorrha­ ging with or without hydrops, and gall bladder agenesis. Male and female Sprague Dawley rats were exposed to methyl bromide by whole-body inhalation exposure, 6 h/day, 5 days/week at concentrations of 0, 3, 30, or 90 ppm in a two-generation reproduction study (American Biogenics Corporation, 1986). Two litters were produced for each generation. No deaths or noteworthy antemortem clinical findings were observed over the course of the study. The 90-ppm F0 males had significantly decreased body weights at five of the 10 premating intervals and at final sacrifice. No other decreases in body weights were observed among the F0 generation or during the F1 generation prior to the gestational period for the F2a litter. A slight depression of body weight was noted during the gestation and lactation periods for the 90-ppm F1 dams. Reproductive performance was not altered by methyl bromide exposure and there were no significant differences in pup survival. No methyl bromide-related anomalies were noted for the progeny. Gross pathologic examination revealed no treatment-related lesions in either the parental animals or their progeny. Mean brain weight for the 90-ppm males (P0 and F1) and females (F1) were decreased. Increased liver-tobody weight ratio for the 90-ppm P0 males and females and increased heart-to-brain weight ratios for the 90-ppm F1 females were noted. No other significant differences were seen in the parent organ weight data. No significant differences in the F1b progeny body and organ weights were noted, but statistically significant decreases in final body weights were observed for the 90-ppm F2b males and females and the 30-ppm F2b females. F2b progeny organ weights were significantly reduced for 90-ppm female

Chapter | 105  Methyl Bromide

brain, heart, kidney, and liver weights, the 30-ppm female liver weight, and the 30- and 90-ppm female liver-to-brain weight ratio. The 30- and 90-ppm F2b female brain-to-body weight ratio was increased. There were no other significant differences noted for progeny. Microscopic examination of the reproductive organs and abnormal tissues revealed no treatment-related lesions.

105.3.5  Chronic Toxicity and Oncogenicity: Inhalation Wistar rats were exposed (whole body) to methyl bromide at atmospheric concentrations of 0, 3, 30, or 90 ppm on a 6 h/day, 5 days/week basis for 29 months (Reuzel et al., 1987, 1991). At the 90-ppm concentration, decreased survival was noted for both males and females from the end of the 2nd year through termination at 29 months. Also, at this concentration, body weights, especially for females, were lower than the control group from week 4 and throughout the remainder of the study, and decreased absolute brain weight was noted for females. No differences in hematology, clinical pathology, or urinalysis were seen at either the 3-month or 1-year intervals. No treatmentrelated evidence of neoplasia was observed in the study. Treatment-related nonneoplastic pathology consisted of an increased incidence of thrombi in the heart and myocardial degeneration for both sexes from the 90-ppm group. Irritation of the nasal cavity characterized by hyperplasia of the olfactory epithelium was seen in a time-related fashion for all methyl bromide-treated groups. Dose-dependent increases in the incidences of degenerative and hyperplastic changes of the nasal olfactory epithelium were observed. The lesions were characterized as very slight, slight, or moderate. A statistically significant increase was found between controls and the low-dose group (3 ppm) at the end of the exposure period (29 months). However, frequency of this lesion also increased in the controls (age dependence) from 12–24 months to 29 months of age. In addition, all but one of the lesions in the 3-ppm exposure level group were described as slight or very slight. Moreover, one moderate lesion of the nasal mucosa was also observed in a control animal at the 24-month sacrifice interval. The NOEL for this lesion was 90 ppm after 12 months of exposure, 3 ppm after 24 months of exposure, and 3 ppm after 29 months of exposure. The Gotoh et al. (1994) study directly correlates with the Reuzel et al. (1987, 1991) study at the 24-month interval. Gotoh et al. exposed F344 rats to methyl bromide concentrations of 0, 4, 20, and 100 ppm on a 6 h/day, 5 day/week basis for 104 weeks. After 24 months of expos­ure, increased incidences of necrosis and respiratory metaplasia of the olfactory epithelium were seen for male rats exposed to 100 ppm methyl bromide; these findings were marginally increased for the female rats at 100 ppm.

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Metaplasia was noted for 22% of control males and 6% of control females at the 24-month terminal interval. These results show that metaplasia produced in the rat olfactory epithelium was a threshold response upon chronic inhal­ ation exposure to methyl bromide and that a high control incidence of metaplasia is noted in aged rats. B6C3F1 mice were exposed via inhalation (whole body) to concentrations of 0, 10, 33, or 100 ppm methyl bromide on a 6 h/day, 5 days/week basis for 2 years (NTP, 1992). The 100-ppm exposure concentration clearly exceeded an acceptable maximum tolerated dose for car­cinogenicity testing. This exposure concentration was terminated after 20 weeks due to debilitating neurotoxicity and mortalities; these animals were exposed to untreated air for the remainder of the 2-year study period. Interim sacrifice of 10 mice/sex/treatment level was performed after 6 and 15 months of exposure. Neurobehavioral testing was performed on selected animals every 3 months. Clinical signs of neurotoxicity, consisting of tremors, paralysis, gait disturbances, and abnormal posture, were noted for 100-ppm males (78%) and females (43%). Similar findings were seen for a few (2–3%) of the 33-ppm-exposed animals. After 3 months of exposure, neurobehavioral changes were noted for 100-ppm males and females. Neurobehavioral testing also revealed changes in the 10- and 33-ppm groups after 6 months of exposure. Decreased body weights were observed in females dosed at 33 ppm and in both sexes dosed at 100 ppm. Exposurerelated histologic changes were generally limited to the 100-ppm animals and consisted of findings in the brain (degeneration of the cerebrum and cerebellum), heart (degeneration and cardiomyopathy), sternal dysplasia, and either necrosis or metaplasia of the olfactory epithelium.

105.3.6  Chronic Toxicity and Oncogenicity: Dietary Methyl bromide was evaluated for chronic toxicity and oncogenicity in a 24-month dietary toxicity study (Mertens, 1997) in Sprague-Dawley rats. Because of the volatile nature of methyl bromide and feeding characteristics of rats, it was not possible to conduct the study using feed fumigated with methyl bromide. For purposes of this study, methyl bromide was microencapsulated and mixed into the rodent diet. Methyl bromide dietary concentrations were 0.5, 2.5, 50, and 250 ppm (0.02, 0.11, 2.20, and 11.10 mg/kg/day for males and 0.3, 0.15, 2.92, and 15.12 mg/kg/day for females, respectively). Basal diet and placebo (microcapsules without methyl bromide) control groups were treated on a comparable regimen. No methyl bromiderelated effects were seen on survival, clinical condition, hematology, serum chemistry, urinalysis, organ weights, ophthalmologic assessments, or macroscopic and microscopic pathology evaluations. Food consumption, mean

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body weights, and mean body weight gains were reduced in the 250-ppm males and females during the rapid growth phase for animals during the first 12–18 months of the study. During the first 18 months of the study, mean body weight gain for males were 9–21% lower than male control groups, while mean body weight gain for females was 7–22% lower than female control groups. Typical of chronic toxicity studies, food consumption, and body weight gains during the 2nd year of the study were comparable to controls as the mature animals reached adult body weight plateau. No evidence of oncogenicity was seen in this study. In a 12-month dietary safety study (Newton, 1995), beagle dogs were exposed to methyl bromide fumigated feed at dietary concentrations of 0, 0.5, 1.5, or 5 ppm (0, 0.06, 0.13, and 0.27 mg/kg/day for males and 0, 0.07, 0.12, and 0.26 mg/kg/day for females, respectively). Prestudy trials were conducted to determine methyl bromide fumigation concentrations and postfumigation intervals required to achieve desired concentrations over a 1-h feeding period. No toxicologically significant methyl bromide effects were seen in clinical observations, body weight, body weight gain, food consumption, clinical pathology, urinalysis, ophthalmology, absolute or relative organ weights, and macroscopic or microscopic pathology. Based on the results of this study, the NOEL for methyl bromide when administered via fumigated feed to beagle dogs was greater than 5 ppm (0.27 mg/kg/day for males and 0.26 mg/kg/day for females).

105.3.7  Neurotoxicity In an acute neurotoxicity study (Driscoll and Hurley, 1993), male and female CD (Sprague-Dawley) rats were exposed via inhalation for 6 h to methyl bromide at concentrations of 0, 30, 100, or 350 ppm. Animals were assessed for clin­ ical signs and changes in body weights. Neurobehavioral evaluations were performed within 3 h of exposure and at 7 and 14 days postexposure. These evaluations included the functional observation battery and motor activity assessments. After 15 days, animals were euthanized, necropsied, and examined for gross pathologic changes, and brains were weighed. In addition, microscopic evaluations were performed on central and peripheral nervous tissue. All animals survived to study termination. No methyl ­bromide-induced effects were noted for body or brain weights. Neurobehavioral effects were observed only in the 350-ppm-exposed group and were limited to the 3-h postexposure assessment. Effects noted in male and female rats consisted of decreased arousal, increased incidences of drooping or half-shut eyelids, piloerection, decreased rearing, depressed body temperature, and markedly decreased motor activity. The 350-ppm males had a decreased tail pinch response, while females from this group showed increased urination and abnormal air righting response. No

Hayes’ Handbook of Pesticide Toxicology

treatment-related histological findings were seen in nervous system or nasal tissues. In a subchronic inhalation neurotoxicity study (Norris et al., 1993), CD (Sprague-Dawley) rats were exposed to methyl bromide concentrations of 0, 30, 70, or 140 ppm. Exposure was 6 h/day, 5 days/week for 13 weeks. At the 140-ppm concentration, two male rats died during the 1st month. Clinical signs observed for these rats included convulsions, tremors, hyperactivity, rapid respiration, and salivation. Mean body weights were significantly lower than the controls. Neurologic evaluations for males revealed increased hind limb splay (weeks 4, 8, 13), abnormal air righting reflex (week 13), and decreased forelimb grip strength (week 13). Female rats demonstrated lower arousal scores (weeks 8, 13), decreased rearing (weeks 4, 8, 13), and significantly decreased motor activity (week 13). Mean absolute brain weights were significantly lower for both sexes. No differences were noted for the relative brain weights, indicating that lower absolute brain weight was a reflection of the generally lower body weights for treated animals. Gross lesions were limited to moderate to severe brain hemorrhage in the two 140-ppm male animals that died. Microscopic lesions in the brain were found in these two males and in one 140-ppm male that survived the 13week exposure. Microscopic lesions in the brain were seen in these three males, and consisted of neuronal necrosis in the hippocampus, necrosis and malacia in the cerebral cortex and basal ganglia, and malacia and/or necrosis in the thalamus and midbrain. The lesions were more severe for one male that was found dead and one male that survived the 13-week exposure; both of these animals were noted with convulsions during the study, suggesting that some of the microscopic brain findings may have been secondary effects of brain swelling related to the convulsions. One additional 140-ppm male had slight neuronal edema in the hippocampus. Other lesions in the 140-ppm group consisted of minimal regenerative dysplasia of the olfactory epithelium of the nasal cavity in three males and three females, and minimal peripheral nerve degeneration in two males and two females. In the 70-ppm group, lower mean body weights and weight gain were seen for females from week 9 onward of the study. Neurologic findings were limited to slightly decreased forelimb grip strength (week 13) in males and decreased motor activity (week 13) in females. Although the mean absolute brain weight for females was statistically significantly decreased (5% lower than the control group), no difference was seen for the relative brain weight, and no microscopic pathology was seen. At 30 ppm, the mean absolute brain weight for females was statistically significantly lower than control (5%); however, no difference was seen in relative brain weight, and no microscopic pathology in the brain was seen. Peripheral nerve degeneration was observed in one female rat. This finding was considered incidental since nerve degeneration was not seen in animals from the 70-ppm group.

Chapter | 105  Methyl Bromide

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A contemporaneous developmental neurotoxicity study via the inhalation route was conducted with methyl bromide (Beck, 2005). In that study, time-mated female Sprague-Dawley rats were exposed for 6 h/day to 0, 5, 25, or 50 ppm methyl bromide from gestation day 6 through 20. During lactation days 5–20, the females were exposed to the appropriate atmospheres simultaneously with their litters in modified inhalation cylinders for 6 h/day. There was no test article-related F0 mortality and no evidence of maternal toxicity at any exposure level. For the F1 progeny, survival during the preweaning and postweaning periods, mean birth weight, and clinical condition was unaffected at all exposure levels. Mean body weight gain in 50-ppm group progeny were statistically significantly decreased from the controls from postnatal days 4–7 through days 14–17, with body weights 8–10% lower than controls on postnatal days 13, 17, and 21. Lower body weights continued through postnatal days 60 and 49 in males and females, respectively. Males and females in the 50-ppm group exhibited delayed day of acquisition of balanopreputial separation and vaginal patency, slightly reduced forelimb grip strength, and a significant reduction in total and ambulatory activity levels during the postnatal day 21 locomotor activity assessment. However, no effect on total or ambulatory activity was observed at postnatal day 61. Despite the reduction in activity levels, the normal pattern of habituation was noted for all exposure groups. No effects were seen at any methyl bromide concentration for functional observational battery assessments, startle response, ability to swim, or learning and memory capacity at any age.

90 or 200 ppm. As discussed previously, severe effects were seen in sustentacular and mature sensory cells, while basal cells remained intact. Regeneration of the olfactory epithelium was seen as early as day 3 of exposure despite continued exposure at these high methyl bromide concentrations. The recovery of the olfactory epithelium was essentially complete by 10 weeks after exposure. The rapid recovery would be expected since the nonaffected basal cells regenerate sensory and sustentacular cells. Olfactory function, as measured by food finding activity, was impaired in ­animals exposed to 200 ppm only. Recovery of this function was evident by 4–6 days postexposure, which is earlier than the time course for histological recovery. In another study (Hastings et al., 1994), morphologic and biochemical (carnosine content of the olfactory bulb, a biomarker for integrity of the olfactory epithelium) evaluations were conducted to further explore methyl bromide exposure effects and recovery. Prior to treatment, rats were food-deprived and trained to find buried food pellets as a measure of olfactory function. The rats were exposed to a methyl bromide concentration of 200 ppm on a 4-h/day, 4 days/week, 2-week regimen. After a single exposure, extensive damage to the olfactory epithelium, reduced carnosine content, and impaired olfactory function were observed. Although exposure continued, olfactory function began to improve after the first exposure. This recovery proceeded even though persistent thinning and disorgan­ ization of the olfactory epithelium and decreased carnosine levels in the olfactory bulb were present. Regeneration of the olfactory epithelium was complete approximately 30–40 days after the last exposure.

105.3.8  Specific Target Organ Effects

105.4  Metabolism

Methyl bromide is an unusual respiratory toxicant in the rat because methyl bromide is specifically toxic to the olfactory epithelium, while other nasal epithelia are unaffected (Hurtt et al., 1988). Within the olfactory epithelium, methyl bromide only affects specific cell types. The major components of the rat olfactory epithelium are basal cells, long ducts of Bowman’s glands, sensory cells, and sustentacular or support cells. Studies using histochemical techniques clearly showed that methyl bromide specifically induces degeneration of sensory and sustentacular cells, while sparing basal cells from which sensory and sustentacular cells are regenerated. Hurtt et al. (1988) evaluated the time course for regeneration of the olfactory epithelium following short-term expos­ ures to methyl bromide. Male rats were exposed to 200 ppm methyl bromide for 6 h/day for 1–5 days. Air-exposed animals served as controls. In a companion study, animals were exposed to 0, 90, or 200 ppm for 6 h, and olfactory function was assessed by the ability of food-deprived animals to locate buried food pellets. Destruction of the olfactory epithelium was evident after a single 6-h exposure to

Inhalation was the primary route of exposure for the majority of the methyl bromide toxicology studies, and is the most probable route of exposure for humans. As a result, the metabolism of methyl bromide has been almost exclusively evaluated by inhalation.

105.4.1  Absorption Medinsky et al. (1984, 1985) evaluated the uptake of methyl bromide upon 6-h exposure of rats to concentrations of 1.6, 9.0, 170, or 310 ppm. Methyl bromide uptake was found to be linear over all exposure concentrations with the exception of the 310-ppm concentration. It was noted that suspected nasal irritation at 310 ppm may have reduced the total amount of methyl bromide inhaled due to decreased tidal and minute volumes (Medinsky et al., 1985). Total methyl bromide absorbed was 9 or 40 mol/kg body weight after exposure to 1.6 ppm (50 nmol/l) or 9.0 ppm (300 nmol/l), respectively. Uptake at the lower levels was approximately 48%. Uptake at 5700 (170 ppm) and

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10,400 (310 ppm) nmol/l was 37% and 27%, respectively (Medinsky et al., 1985). Andersen et al. (1980) found methyl bromide to exhibit rapid, first-order uptake kinetics. Saturation was not reached until concentrations causing animal death were achieved. Gargas and Andersen (1982) also demonstrated that methyl bromide uptake and metabolism followed first order over a broad range of exposure concentration (100–3000 ppm). Honma et al. (1985) also found methyl bromide to be rapidly absorbed and distributed. Methyl bromide concentrations in blood, liver, adipose, and brain reached maximum levels within 1 h after the start of ex­posure and maintained almost the same levels during the 8-h exposure. Equilibrium between methyl bromide air and tissue concentrations occurred rapidly.

105.4.2  Distribution Methyl bromide is rapidly and widely distributed in tissues immediately after exposure. Bond et al. (1985) investigated the tissue distribution of 14 C methyl bromide. Radioactivity was widely distributed in tissues immediately following exposure with highest levels of 14C found in lung, adrenal, kidney, liver, and nasal turbinates. Low concentrations of 14C were also detected in other tissues. The liver was the only tissue examined immediately after exposure that contained a large percentage (approximately 17%) of the absorbed methyl bromide. Approximately 80% of the initial amount of tissue radiolabel was eliminated by 65 h. Elimination half-lives varied from 1.5 to 8 h, with the exception of the liver with an elimination half-life of 33 h. Using nonradiolabeled methyl bromide, Honma et al. (1985) investigated the distribution of methyl bromide into liver, fat, brain, muscle, kidney, and blood upon inhal­ ation exposure of rats to 250 ppm for 8 h. Methyl bromide concentrations in all tissues listed reached maximum levels within 1 h after the start of the exposure and was found at approximately the same tissue concentration through the remainder of the exposure period. The highest tissue concentration was found in fat. Methyl bromide was rapidly eliminated from rat tissues following the cessation of exposure, with a half-life of approximately 30 min in the early postexposure period. At 48 h postexposure, methyl bromide was not detected in any tissue examined. Medinsky et al. (1985) evaluated tissue distribution of methyl bromide 66 h after exposure to a range of vapor concentrations. Significant levels of 14C (approximately 20% of the absorbed methyl bromide dose) remained in tissues 66 h after exposure. The highest level (approximately 20% of the 14C) was associated with the liver. Other tissues having appreciable concentrations of 14C included the lungs, nasal turbinates, and kidneys. Very low concentrations of 14C (less than 1 mol of methyl bromide equivalents/g of tissue) were found in nervous tissues (spinal cord and brain).

Hayes’ Handbook of Pesticide Toxicology

Tissue disposition after oral or intraperitoneal (IP) injection of 250 mol/kg 14C-methyl bromide was also evaluated (Medinsky et al., 1984). Approximately 14 and 17% of the radioactivity administered was found in the tissues and carcass 72 h after oral and IP dosing, respectively. Analysis of individual tissues indicated liver was the major organ for retention of radioactivity after administration of methyl bromide. Other tissues containing significant amounts of radioactivity (10 nmol or 1% of the total dose) included kidneys, testes, lung, heart, stomach, and spleen. As expected, the only tissue showing significantly higher radioactivity level after oral administration, as compared with IP injection, was the stomach.

105.4.3  Identification and Quantitation of Metabolites In all methyl bromide metabolism studies, carbon dioxide was the major metabolite. Approximately 47% of the total methyl bromide dose was excreted as 14CO2 in expired air (Bond et al., 1985). About 1% of the total 14Cmethyl bromide absorbed was exhaled as 14C-methyl bromide. Smaller quantities of the radiolabeled material were excreted in the urine and feces, with about 22% and 2% of the total absorbed 14C-methyl bromide excreted by these routes, respectively. All radioactivity in excreta was identified as methyl bromide degradates/metabolites. No evidence of parent chemical was found in any of the excreta samples (Bond et al., 1985). Gargas and Andersen (1982) showed that bromine ion is released in the initial metabolism of a variety of brominated hydrocarbons. The study results showed that bromine ion is retained in extracellular fluid, is stable to further biotransformation, and is very slowly excreted. The first-order rate constant for the release of bromine from methyl bromide is 0.32/kg/h. The elimination of bromine from rat tissue is slower than that of methyl bromide (Honma et al., 1985). Peak concentrations of bromine in blood, kidneys, and liver occurred 4–8 h after methyl bromide exposure, and the half-life of bromine in these tissues was approximately 5 days. There was no correlation between the duration of bromine retention and observed signs of neurotoxicity (Gargas and Andersen, 1982). In a methyl bromide inhalation study in mice, Alexeeff and Kilgore (1985) followed mouse tissue bromine concentrations through 7 days postexposure. No bromide ion was detected in any tissue 1 week after exposure to concentrations up to 2.72 mg/l air. Over 95% of the bromide ion in exposed mice was eliminated within 2.5 days. The bromide ion levels were highest in liver and kidney, and lowest in whole blood. Lung and brain bromide ion levels were intermediate. Honma et al. (1985) also investigated bromine and methanol concentrations in response to treatment with methyl bromide. Peak concentrations of bromine in blood, kidneys, and liver occurred 4–8 h after methyl bromide

Chapter | 105  Methyl Bromide

exposure, and the half-life of bromine in these tissues was about 5 days. Methanol production was not significant.

105.4.4  Excretion Excretion of methyl bromide is primarily as exhaled CO2. Very low levels of parent methyl bromide are found in expired air. Greater than 85% of the total amount of 14C exhaled as CO2 and excreted in urine and feces was elim­ inated within 24 h (Bond et al., 1985). CO2 excretion exhibited a biphasic elimination pattern with 85% of the 14CO2 excreted with a half-life of approximately 4 h, and 15% excreted with a half-life of approximately 11 h. The elim­ ination half-life of 14C in urine and in feces was approximately 10 h and 16 h, respectively. Analysis of excreta samples in which there was sufficient radioactivity showed no evidence of parent methyl bromide.

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f­ umigation. Since human exposure is more likely to occur by the inhalation route, methyl bromide has been extensively evaluated in a number of toxicologic studies ranging from acute through chronic exposures. Neurotoxicity was the primary effect following inhalation exposures in acute and subchronic neurotoxicity studies in rats, in a series of specifically designed neurotoxi­city studies in dogs, and in a developmental neurotoxicity study with direct inhalation exposure of dams and pups. Neurobehavioral findings and neuropathologic changes were generally seen at high-exposure concentrations in these studies. Collectively, the dog inhalation studies showed that neurologic effects were related to both concentration and number of exposures. In addition to neuro­ toxicity, methyl bromide inhalation exposures lead to degeneration of the nasal olfactory epithelium. The nasal findings are reversible and regeneration has been noted during continued exposure.

105.5  Human exposure The potential routes of human exposure to methyl bromide are oral (through consumption of fumigated food products), dermal (skin contact), or inhalation (exposure to methyl bromide gas). Extensive studies have shown residues of methyl bromide found in crops grown on fumigated soils are virtually undetectable. In addition, methyl bromide concentrations in commodities treated postharvest decrease rapidly after required aeration, and are undetectable after relatively short periods of time. Tolerance levels for the metabolite of methyl bromide (inorganic bromide) in treated foods have been established by the U.S. EPA. These tolerances further ensure humans are not exposed to unsafe levels of methyl bromide’s metabolite in foods. In summation, there is no significant likelihood of oral exposure to methyl bromide through consumption of treated food. Human exposure is more likely to occur by inhal­ation. People living in close proximity to fumigated fields, greenhouses, or structures are protected from the risk of significant inhalation exposure through special notice requirements, safety precautions, and the use of buffer zones. The potential for dermal or inhalation exposure to methyl bromide is highest for applicators and other personnel involved in manufacturing, filling, handling, or applying methyl bromide. Strictly applied safety measures in manufacturing and filling installations limit the potential risk of exposure to plant personnel. In addition, fumigators/applicators are protected from dermal and inhalation exposures through adherence to strict safety procedures and the use of protective equipment.

Conclusion Methyl bromide is primarily used for soil fumigation, commodity/quarantine treatment, and structural

References Alexeeff, G. V., and Kilgore, W. W. (1983). Methyl bromide. Residue Rev. 88, 101–153. Alexeeff, G. V., and Kilgore, W. W. (1985). Determination of acute toxic effects in mice following exposure to methyl bromide. J. Toxicol. Environ. Health 15, 109–123. American Biogenics Corporation (1986). Two-Generation Reproduction Study via Inhalation in Albino Rats Using Methyl Bromide. Unpublished report from Study 450–1525. Andersen, M., Gargas, M., Jones, R., and Jenkins, L. (1980). Determination of the kinetic constants for metabolism of inhalated toxicants in vivo using gas uptake measurements. Toxicol. Appl. Pharmacol. 54, 100–116. Anger, W. K., Setzer, J. V., Russo, J. M., Brightwell, W. S., Wait, R. G., and Johnson, B. L. (1981). Neurobehavioral evaluation of soil and structural fumigators using methyl bromide and sulfuryl fluoride. Neuro Toxicology 7, 137–156. Araki, A., Kato, F., Matsushima, T., Ikawa, N., and Nozaki, K. (1995). Methyl bromide—micronuclei induction of methyl bromide in rats and mice by subchronic inhalation test. Environ. Mut. Commun. 17, 47–56. ATSDR (1991). Toxicological Profile for Bromomethane. U.S. Department of Health and Human Services, Public Health Services Agency for Toxic Substances and Disease Registry, Publication 91–06. Bakhishev, G. N. (1973). Relative toxicity of aliphatic halohydrocarbons to rats. Farmakol. Toksikol. 8, 140–142 [In Russian]. Balander, P. A., and Polyak, M. G. (1962). Toxicological characteristics of methyl bromide. J. Gig. I. Toksikol. 60, 412–419. Beck, M. J. (2005). An inhalation developmental neurotoxicity study of methyl bromide in rats. Unpublished report from WIL Research Laboratories, Inc. Project WIL-186039. Bentley, K. S. (1994). Detection of Single Strand Breaks in Rat Testicular DNA by Alkaline Elution Following In Vivo Inhalation Exposure to Methyl Bromide.” Unpublished report from E. I. DuPont Haskell Laboratories, Project 9714–001: MBIP/21/ALK/ HASK: 999.

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Bond, J. A., Dutcher, J. S., Medinsky, M. A., Henderson, R. F., and Birnbaum, L. S. (1985). Disposition of [14C]methyl bromide in rats after inhalation. Toxicol. Appl. Pharmacol. 78, 259–267. Breslin, W. J., Zablotny, C. L., Bradley, G. J., Nitschke, K. D., and Lomax, L. G. (1990a). Methyl Bromide Inhalation Teratology Probe Study in New Zealand White Rabbits. Unpublished study from Dow Chemical Company Toxicology Laboratory. Breslin, W. J., Zablotny, C. L., Bradley, G. J., and Lomax, L. G. (1990b). Methyl Bromide Inhalation Teratology Study in New Zealand White Rabbits. Unpublished study from Dow Chemical Company Toxicology Laboratory. Danse, L. H. J. C., van Velsen, F. L., and van der Heijden, C. A. (1984). Methyl bromide: Carcinogenic effect in the rat forestomach. Toxicol. Appl. Pharmacol. 72, 262–271. Djalali-Behzad, G., Hussain, S., Ostermann-Golker, S., and Segerbaeck, D. (1981). Estimation of genetic risks of alkylating agents. VI. Exposure of mice and bacteria to methyl bromide. Mutat. Res. 84, 1–9. Driscoll, C. D., and Hurley, J. M. (1993). Methyl Bromide: Single Exposure Vapor Inhalation Neurotoxicity Study in Rats. Unpublished report from Bushy Run Research Center, Project 92N1197. Eustis, S. L., Haber, S. B., Drew, R. T., and Yang, R. S. H. (1988). Toxicology and pathology of methyl bromide in F344 rats and B6C3F1 mice following repeated inhalation exposure. Fundam. Appl. Toxicol. 11, 594–610. Gargas, M., and Andersen, M. (1982). Metabolism of inhaled brominated hydrocarbons: validation of gas uptake results by determination of stable metabolite. Toxicol. Appl. Pharmacol. 66, 55–68. Garry, V. F., Nelson, R. L., Griffith, J., and Harkins, M. (1990). Preparation of human study of pesticide applicators: sister chromatid exchanges and chromosomal aberrations in cultured human lymphocytes exposed to selected fumigants. Teratolog. Carcinog. Mutagen. 10, 21–29. Gotoh, K., Nishizawa, T., Yamaguchi, T., Kanou, H., Kasai, T., Ohsawa, M., Ohbayyashi, H., Aiso, S., Ikawa, N., Yamamoto, S., Noguchi, T., Nagano, K., Enomoto, M., Nozaki, K., and Sakabe, H. (1994). Two year toxicological and carcinogenesis studies of methyl bromide in F344 rats and BDF1 mice—Inhalation studies. In “Proceedings: Second Asia-Pacific Symposium on Environmental and Occupational Health.” Haber, S. B., Drew, R. T., Eustis, S., and Yang, R. S. H. (1985). Methyl bromide toxicity: a target organ? Toxicologist 5, 130 (Abstract 518). Hastings, L., Andringa, A., and Miller, M. A. (1994). Exposure of the olfactory system to toxic compounds: structural and functional consequences. Inh. Toxicol. 6, 437–440. Hine, C. H. (1969). Methyl bromide poisoning: A review of ten cases. J. Occup. Med. 11, 1–10. Honma, T., Miyagawa, M., Sato, M., and Hasegawa, H. (1985). Neurotoxicity and metabolism of methyl bromide in rats. Toxicol. Appl. Pharmacol. 81, 183–191. Hurtt, M. E., Morgan, K. T., and Working, P. K. (1987). Histopathology of acute toxic responses to selected tissues from rats exposed by inhalation to methyl bromide. Fundam. Appl. Toxicol. 9, 352–365. Hurtt, M. E., Thomas, D. A., Working, P. K., Monticello, T. M., and Morgan, K. T. (1988). Degeneration and regeneration of the olfactory epithelium following inhalation exposure to methyl bromide: Pathology, cell kinetics and olfactory function. Toxicol. Appl. Pharmacol. 94, 311–328. Irish, D. D., Adams, E. M., Spencer, H. C., and Rowe, V. K. (1940). The response attending exposure of laboratory animals to vapors of methyl bromide. J. Ind. Hyg. Toxicol. 22, 218–230.

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Japanese Ministry of Labour (1992). Toxicology and Carcinogenesis Studies of Methyl Bromide in F344 Rats and BDF Mice (Inhalation Studies). Unpublished report from the Industrial Safety and Health Association, Japanese Bioassay Laboratory, Tokyo. Kato, N., Morinobu, S., and Ishizu, S. (1986). Subacute inhalation experiment for methyl bromide in rats. Ind. Health 24, 87–103. Katz, A. J. (1985). Genotoxicity of methyl bromide in somatic cells of Drosophila larvae. Environ. Mutagen. 7, 13. Katz, A. J. (1987). Inhalation of methyl bromide gas induces mitotic recombination in somatic cells of Drosophila melanogaster. Mutat. Res. 192, 131–135. Kiplinger, G. A. (1994). Methyl Bromide: Acute Oral Toxicity Comparison Study of Microencapsulated Methyl Bromide and Liquid Methyl Bromide in Albino Rats. Unpublished report from WIL Research Laboratories, Project WIL-49011. Kramers, P. G. N., Bissumbhar, B., and Mout, H. C. A. (1985a). Studies with gaseous mutagens in Drosophila melanogaster. In “Short Term Bioassays in the Analysis of Complex Environmental Mixtures IV” (M. D. Waters, S. S. Sandhu, J. Lewtas, L. Claxton, G. Straus, and S. Nesnow, eds.), pp. 65–73. Plenum, New York/London. Kramers, P. G. N., Voogd, C. E., Knaap, A. G. A. C., and Van der Heijden, C. A. (1985b). Mutagenicity of methyl bromide in a series of short term assays. Mutat. Res. 155, 41–47. McGregor, D. B. (1981). Tier II Mutagenic Screening of 13 NIOSH Priority Compounds. Report 32. Individual Compound Report: Methyl Bromide. National Institute of Occupational Safety and Health, Cincinnati, OH, PB83-130211. Medinsky, M., Bond, J., Dutcher, J., and Birnbaum, L. (1984). Disposition of [14C]-methyl bromide in rats after inhalation. Toxicology 32, 187–196. Medinsky, M., Dutcher, J., Bond, J., Henderson, R., Mauderly, J., Snipes, M., Mewhinney, J., Cheng, Y., and Birnbaum, L. (1985). Uptake and excretion of [14C]-methyl bromide as influenced by exposure concentration. Toxicol. Appl. Pharmacol. 78, 215–225. Mertens, J. J. W. M. (1997). A 24-Month Chronic Dietary Study of Methyl Bromide in Rats. Unpublished report from WIL Research Laboratories, Project WIL-49014. Moriya, M., Ohta, T., Watanabe, K., Miyazawa, T., Kato, K., and Shirasu, Y. (1983). Further mutagenicity studies on pesticides in bacterial reversion assay systems. Mutat. Res. 116, 185–216. Naas, D. J. (1990). Acute Oral Toxicity Study in Beagle Dogs with Methyl Bromide. Unpublished report from WIL Research Laboratories, Inc., Project WIL-49006. Newton, P. E. (1994a). A Four Week Inhalation Toxicity Study of Methyl Bromide in the Dog. Unpublished report from Pharmaco LSR, Project 93–6068. Newton, P. E. (1994b). An Up-and-Down Acute Inhalation Toxicity Study of Methyl Bromide in the Dog. Unpublished report from PharmacoLSR, Project 93–6067. Newton, P. E. (1995). A Chronic (12-Month) Toxicity Study of Methyl Bromide Fumigated Feed in the Dog. Unpublished report from Pharmaco-LSR, Project 94–3186. Norris, J. C., Driscoll, C. D., Hurley, and J. M. (1993). Methyl Bromide: Ninety-Day Vapor Inhalation Neurotoxicity Study in CD Rats. Unpublished report from Bushy Run Research Center, Project 92N1172. NTP (1992). Toxicology and Carcinogenesis Studies of Methyl Bromide (CAS No. 74–83–9) in B6C3F1 Mice (Inhalation Studies). National Toxicology Program Technical Report Series 385.

Chapter | 105  Methyl Bromide

Ong, J. M., Stewart, J., Wen, Y., and Whong, W. (1987). Application of SOS umu-test for the detection of genotoxic volatile chemicals and air pollutants. Environ. Mutagen. 9, 171–176. Reuzel, P. G. J., Kuper, C. F., Dreef van der Meulen, H. C., and Hollanders, V. M. H. (1987). Chronic (29-Month) Inhalation Toxicity and Carcinogenicity Study of Methyl Bromide in Rats. Unpublished Report From CIVO Institutes TNO. Reuzel, P. G. J., Dreef van der Meulen, H. C., Hollanders, V. M. H., Kuper, C. F., Feron, V. J., and van der Heijden, C. A. (1991). Chronic inhalation toxicity and carcinogenicity study of methyl bromide in Wistar rats. Food Chem. Toxicol. 29, 31–39. Schaefer, G. J., Kirkpatrick, D. T., Holson, J. F., Chengelis, C. P., Regan, K. S., and Piccirillo, V. J. (2002). A 6-week inhalation toxicity study of methyl bromide in dogs. Unpublished report from WIL Research Laboratories, Inc., Project WIL-440001. Sikov, M. R., Cannon, W. C., Carr D. B., Miller, R. A., Montgomery, L. F., and Phelps, D. W. (1981). Teratologic Assessment of Butylene Oxide, Styrene Oxide and Methyl Bromide. Battelle Pacific Northwest Laboratories, Contract 210–78–0025, Division of Biomedical and Behavioral Science, National Institute for Occupational Safety and Health, U. S. Department of Health and Human Services. Simmon, V. F., Kauhanen, K., and Tardiff, R. G. (1977). Mutagenic activity of chemicals identified in drinking water. In “Progress in Genetic

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Toxicology” (D. Scott, B. A. Bridges, and F. M. Sobels, eds.), pp. 249–258. Elsevier/North-Holland Biomedical Press, Amsterdam. Tucker, J. D., Xu, J., Stewart, J., Baciu, P. C., and Ong, T. (1986). Detection of sister chromatid exchanges induced by volatile genotoxicants. Teratog. Carcinog. Mutagen. 6, 14–21. WHO (1995). Environmental Health Criteria 166, Methyl Bromide. Published under the joint sponsorship of the United Nations Environment Programme, the International Labour Organisation, and the World Health Organization, Geneva. Wilmer, J. W. G. M., Reuzel, P. G. J., and Dreef van der Muelen, H. C. (1983). Subchronic (13 week) inhalation toxicity study of methyl bromide in rats. CIVO Rep. V82.378. CIVO Institutes, TNO, Zeist, The Netherlands. Yamano, Y. (1991). Experimental study on methyl bromide poisoning in mice. Acute inhalation study and the effects of glutathione as an antidote. Jpn. J. Ind. Health 33, 23–30 [in Japanese]. Zwart, A. (1988). Acute Inhalation Study of Methyl Bromide in Rats. CIVO Rep. V88. 127/27. CIVO Institutes, TNO, Zeist, The Netherlands. Zwart, A., Arts, J. H. E., Ten Berge, W. F., and Appelman, L. M. (1992). Alternative acute inhalation toxicity testing by determination of the concentration-time-mortality relationship: Experimental comparison with standard LC50 testing. Regul. Toxicol. Pharmacol. 15, 278–290.

Chapter 106

1,3-Dichloropropene W.T. Stott and B.B. Gollapudi Dow Chemical Company, Midland, Michigan

106.1  Chemistry and formulations

106.1.5  Formulations

106.1.1  Chemical Name

Commercial formulations, containing varying amounts of a mixture of cis- and trans-isomers, have historically been marketed under the trademarks Dorlone (admixture with 1,2-dibromoethane), D-D Soil Fumigant, Nemex, Telone, and Vidden D. More recent trademarks include Vorlex (admixture with methylisothiocyanate), Di-Trapex, D-D Super, Telone II Soil Fumigant, and Telone C-17 Soil Fungicide and Nematicide (admixture with chloropicrin). Purified cis-isomer 1,3-dichloropropene has also been marketed under the trademarks Telone-cis and Nematrap. Formulations have typically contained 1–2% of an acid scavenger. Older formulations were often stabilized using epichlorohydrin while epoxidized soybean oil has been utilized in more recent formulations.

1,3-Dichloro-1-propene is the chemical name. The pesticide is typically marketed either as a relatively pure cisisomer or as a mixture of cis- (E) and trans- (Z) isomers.

106.1.2  Structure See Figure 106.1.

106.1.3  Synonyms 1,3-Dichloropropene also is known as -chloroallyl chloride and 1,3-dichloropropylene. The CAS registry number for 1,3-dichloropropene is 542-75-6. The number for the trans-isomer is 10061-02-6; that of the cis-isomer is 10061-01-5.

106.1.4  Physical and Chemical Properties 1,3-Dichloropropene has the empirical formula C3H4Cl2 and a molecular weight of 110.98. It is a white to ambercolored liquid with a sweet, penetrating odor. The density at 25°C is 1.217. The boiling points of the cis- and transisomers are 104 and 112°C, respectively. The flash point is 28°C. The solubility in water at 25°C is approximately 2 g/kg. The compound is miscible with acetone, benzene, carbon tetrachloride, heptane, and methanol.

Cl

Cl Cl

H

H

H cis-

Cl H trans-

Figure 106.1  Structural isomers of 1,3–Dichloropropene. Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

106.2  Uses Introduced in 1945, 1,3-dichloropropene is a soil fumigant nematocide, for preplanting control of parasitic plant nematodes in numerous food and nonfood crops including deciduous fruit and nuts, vines, strawberries, field crops, vegetables, tobacco, tree nurseries, and numerous other specialty crops. Formulated 1,3-dichloropropene is injected into the soil using chisels prior to crop planting at a minimum depth of 10–12 in. below the soil surface. The injected zone is subsequently capped off with soil, which is often then covered with plastic to help maintain concentrations in the soil. Injected 1,3-dichloropropene is believed to volatilize, move through the soil air space, and redissolve into the film of water that surrounds soil particles, where it may exert its toxic effect on soil nematodes. 1,3-Dichloropropene in the soil is lost via chemical hydrolysis in water, metabolism by soil biotica, and evaporation. Efficacy is thus dictated not only by target organism sensitivity, but also by the vapor pressure, diffusion coefficient, distribution in air, water, and soil phases of the soil matrix, and the temperature and moisture content of the treated soil. 2281

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106.3  Hazard identification 106.3.1  Acute Toxicity 1,3-dichloropropene is irritating to eyes and skin of animals. As reviewed by Torkelson (1994), application of mixed isomers of 1,3-dichloropropene to the skin of rabbits up to 4 h (occluded application site) caused a moderate erythema and moderate to severe edema. Mixed isomers of 1,3-dichloropropene also caused a marked redness and slight to moderate chemosis of the conjunctivae immediately following instillation of 0.1 ml into the eyes of rabbits. These effects were gradually reversible and, in most instances, washing with water was effective in averting injury. Similar dermal and ocular effects have been reported for cis-1,3-dichloropropene (Gardner, 1989) and a formulation of 1,3-dichloropropene containing approximately 30% 1,2-dichloropropane (D-D) (Hine et al., 1953). Both mixed isomer 1,3-dichloropropene and cis-1,3-dichloropropene have tested positive in guinea pig skin sensitization assays (Gardner, 1989; Jones, 1988c; Torkelson, 1994). The acute oral, dermal, and inhalation lethality of both mixed isomers of 1,3-dichloropropene and cis-1,3-dichloro­ propene in laboratory animals have also been established (Gardner, 1989; Hine et al., 1953; Jones, 1988a,b; Jones and Collier, 1986a,b; Nitschke et al., 1990a; Torkelson, 1994). The oral LD50 of mixed isomers of 1,3-dichloropropene in rats ranges from 130 to 713 mg/kg in males and 110– 250 to 510 mg/kg in females dependent upon the vehicle and strain used. The oral LD50 for cis-1,3-dichloropropene in rats has been calculated to be 85 to 126 and 117 in males and females, respectively. The dermal LD50 has been reported as greater than 1211 mg/kg mixed isomers in both sexes of rats (unoccluded application site); 1000 and 1300–2000 mg/kg mixed isomers in male and female rats, respectively (occluded application site); 333–540 mg/kg mixed isomers for both sexes of rabbits (occluded application site); and 758–1090 mg/kg cis-1,3-dichloropropene for both sexes of rats (occluded application site). The acute 4-h LC50 value for mixed isomers of 1,3-dichloropropene vapor in rats was 855–1035 ppm for males and 904 ppm for females. Exposed animals had a distinct “garlic” odor and suffered eye and nasal irritation. The acute 4-h LC50 value for cis-1,3-dichloropropene vapor in rats was 670 and 744 ppm for males and females, respectively. Grossly observable eye and respiratory tract irritation was absent following a 2-week observation period. Brief exposure to concentrations in excess of 2700 ppm mixed isomer vapor also caused severe lung, liver, and kidney injury.

106.3.2  Repeated Dose Toxicity The subacute and subchronic toxicity of 1,3-dichloropropene has been examined both orally, via gavage or mixed in feed, and via inhalation. Subacute and subchronic oral toxicity

Hayes’ Handbook of Pesticide Toxicology

studies on relatively modern formulations have been carried out in rats, mice, and dogs by stabilizing the 1,3-dichloropropene by microencapsulation in a starchsucrose matrix and then mixing the encapsulated material into the feed of test animals. Studies reported by Stott et al. (1988) have demonstrated the ready bioavailability of this material once ingested by test animals. Relatively short-term oral dietary toxicity studies employing microencapsulated 1,3-dichloropropene were conducted as a prelude to the subchronic studies summarized below (Haut et al., 1992a,b). Male and female Fischer 344 rats and B6C3F1 mice were administered dosages of 10, 25, 50, or 100 (rats) or 175 (mice) mg/kg/day of microencapsulated 1,3-dichloropropene (mixed isomers) via their diet for 2 weeks. The body weights of both sexes of rats ingesting 50 mg/kg/day and male mice and female mice ingesting 100 and 175 mg/kg/day, respectively, were decreased. Histopathological changes were restricted to rats and consisted of hyperplasia and hyperkeratosis of the nonglandular mucosa of the stomachs of both sexes of rats ingesting 50 mg/kg/day and a single male ingesting 25 mg/kg/day. In a subsequent subchronic rat study, male and female Fischer 344 rats were administered dosages of 5, 15, 50, or 100 mg/kg/day of microencapsulated 1,3-dichloropropene (mixed isomers) via their diet for 13 weeks (Haut et al., 1996). The body weights of males and females ingesting 5 and 15 mg/kg/day, respectively, were decreased. A number of changes in serum biochemical parameters and decreases in organ weights accompanied the depressed body weights of these animals. Histopathological changes were restricted to basal cell hyperplasia and/or hyperkeratosis of the nonglandular mucosa of the stomach of both sexes ingesting 50 mg/kg/day and a single male ingesting 15 mg/kg/day. These changes were at least partially reversible upon ingestion of control feed for 4 weeks. In a subsequent subchronic mouse study, male and female B6C3F1 mice were administered dosages of 15, 50, 100, or 175 mg/kg/day of microencapsulated 1,3-dichloropropene (mixed isomers) via their diet for 13 weeks (Haut et al., 1996). A dose-related decrease in the body weights of males and females ingesting 50 mg/kg/day was observed. Histologic changes consistent with decreased cytoplasmic glycogen and with decreased lipid content were observed in the liver of all treated mice and the kidneys of high-dose group mice. No treatment-related histopathologic effects were reported. Male and female beagle dogs were administered microencapsulated 1,3-dichloropropene via their diets for 13 weeks at concentrations that resulted in mean dosages of 5, 15–16, or 41 mg/kg/day (Stebbins et al., 1999). The body weights of both sexes of dogs were decreased in a dose-related manner relative to controls. The primary effect of 1,3-dichloropropene ingestion was upon erythroid parameters measured in peripheral blood.

Chapter | 106  1,3-Dichloropropene

Calculated erythroid indices and morphologic changes in stained peripheral blood of male and females ingesting 15–16 mg/kg/day indicated the presence of a hypochromic, microcytic anemia. In an early oral toxicity study, male and female Wistar rats were administered dosages of 1, 3, 10, or 30 mg/kg/day of a roughly 78% pure mixed isomer 1,3-dichloropropene formulation 6 days/week for 13 weeks via oral gavage (see Stott et al., 1988; Til et al., 1973). The kidney weights of both sexes of high-dose group animals and males administered 10 mg/kg/day mixed isomer 1,3-dichloropropene were elevated relative to controls. However, no gross or histopathologic changes or alterations in hematologic indices, urinalysis, or serum enzymes accompanied these changes. It was not clear whether toxicity was dictated by 1,3-dichloropropene or of some impurity, possibly 1,2-dichloropropane, which was present at a concentration of nearly 20%. The latter chemical reportedly causes increased liver and kidney weights and hepatic histopathologic changes in rats (Bruckner et al., 1989; IPCS, 1993). The toxicity of inhaled 1,3-dichloropropene in several formulations has also been evaluated over the years. Formulations of 90% purity have been studied in which both sexes of Fischer 344 rats and CD-1 mice were exposed 6 h/day, 5 days/week to mixed-isomer vapor concentrations of 5, 10, or 30 ppm for 4 weeks; 10, 30, or 90 ppm for 13 weeks; and 10, 30, 90, or 150 ppm for 13 weeks (Coate, 1979a,b; Stott et al., 1988). No treatmentrelated effects were observed in rats or mice following a 4-week exposure to up to 30 ppm vapor. However, decreases in body weights were noted and the nasal mucosa and urinary bladder (female mice) were identified as potential target tissues of inhaled 1,3-dichloropropene for 13 weeks. Nasal effects consisted of degeneration of olfactory epithelium and/or hyperplasia of respiratory epithelium in both sexes of rats (30 ppm) and mice (90 ppm) and respiratory metaplasia in olfactory regions of mice (150 ppm). Bladder effects consisted of hyperplasia of the transitional epithelium in female mice only (90 ppm). Subsequent studies were undertaken in which Fischer 344 rats were similarly exposed to cis-1,3-dichloropropene vapor concentrations of 10, 60, or 150 ppm for 2 weeks or 10, 30, or 90 ppm for 13 weeks (Nitschke and Lomax, 1990; Nitschke et al., 1990b). These latter studies also identified body weight changes and/or histopathological changes in the nasal respiratory and olfactory epithelium in both sexes of rats (60–90 ppm). In contrast, in an early inhalation study conducted in 1958, male and female rats, guinea pigs, rabbits, and dogs (females only) were exposed to 1 or 3 ppm mixed isomer 1,3-dichloropropene vapor 7 h/day, 5 days/week, for 6 months (Torkelson and Oyen, 1977). The only effect of exposure that was reported was a “foamy vacuolation or proximal tubule epithelium” in high exposure group male rats. The significance of this change has been questioned in

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view of findings from more recent studies and the diagnosis was characteristic of nephropathy endemic in the sitebred rats used during this early period (Stott et al., 1988). No effects were reported in female rats or other animals exposed to 1 ppm vapor. Exposure of rats and mice to 5, 15, or 50 ppm D-D, an admixture of 1,3-dichloropropene and 1,2-dichloropropene, for 12 weeks revealed liver and kidney weight changes in high-exposure male and female rats, respectively, and diffuse hepatocellular swelling in high-exposure male mice (urinary bladders were not examined) (Parker et al., 1982). As noted, increased liver and kidney weights and hepatic histopathologic changes have been a consistent finding in 1,2-dichloropropene toxicity studies (Bruckner et al., 1989; IPCS, 1993).

106.3.3  Effects on Reproduction Mixed isomer vapors of 1,3-dichloropropene were not embryotoxic or teratogenic in bred rats or inseminated rabbits exposed to 20, 60, or 120 ppm vapors, 6 h/day, during gestation days 6–15 (rats) or 6–18 (rabbits) (Hanley et al., 1987). Maternal toxicity was evidenced at all exposure levels. Exposure of male and female rats to 10, 30, or 90 ppm 1,3dichloropropene vapors for two generations did not adversely affect reproduction or neonatal growth or survival even though 90 ppm proved to be a toxic exposure level (Breslin et al., 1989). Consistent with these results, no treatmentrelated changes in testes weight, sperm count, or sperm morphology occurred in mice 30 days after being injected i.p. with 10, 19, 38, 75, 150, 300, or 600 mg/kg/day 1,3dichloropropene daily, for 5 days (Osterloh et al., 1993). Finally, exposure of male and female rats to 14, 32, or 96 ppm D-D, a low-purity 1,3-dichloropropene formulation containing approximately 30% 1,2-dichloropropane, 6 h/day for 10 weeks did not affect animal mating behavior or fertility (Linnett et al., 1988).

106.3.4  Absorption, Distribution, Metabolism, and Excretion Toxicity and pharmacokinetic data indicate that 1,3-dichloro­ propene is absorbed from the skin, respiratory tract, and gastrointestinal tract. Both inhaled and ingested cis- and trans-isomers were rapidly eliminated from the bloodstream of rats in a biphasic manner consisting of a prominent initial phase with a half-life of approximately 4–7 min followed by a slower phase with a half-life of 22–43 min (Stott and Kastl, 1986; Stott et al., 1998). The predominant routes of excretion of radioactivity in male and female rats and mice following administration of an oral dose of cis, trans, or mixed 1,3-dichloropropene were via the urine (cis, 82–84%; trans, 56–61%; mixed, 51–79%), feces (cis or trans, 2–3%; mixed, 15–21%), and expiration of CO2

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(cis, 2–5%; trans, 23–24%; mixed, 14–18%) (Bartels et al., 2004; Hutson et al., 1971; IPCS, 1993). Excretion and distribution of 1,3-dichloropropene was independent of dose in rats and mice administered up to 50 and 100 mg/kg, respectively (Bartels et al., 2004). In both species, greater than 80% of the administered dosages were excreted within 24 h of dosing. There were no remarkable sex- or speciesrelated differences in excretion routes, kinetics of excretion, or tissue distribution of administered radioactivity in rats (Bartels et al., 2004; Hutson et al., 1971). Humans also rapidly metabolize and excrete inhaled 1,3-D vapor (Waechter et al., 1992). Human volunteers exposed to 1.0 ppm mixed isomer vapor for 6 h absorbed roughly 80% of inhaled 1,3-D. Blood concentrations of 1,3-D rapidly fell following exposure, resulting in no quantifiable concentrations by the first sampling time point postexposure, 10 min, establishing a blood half-life of less than this value. Approximately 90% of the estimated dose of inhaled Telone II was excreted within 36 h of exposure. Dermal absorption of 1,3-D vapor does not appear to be a factor in exposed humans as whole-body uptake has been estimated to be roughly only 2–5% of that absorbed via inhalation (Kezic et al., 1996). The mercapturic acid conjugate of 1,3-dichloropropene and its further oxidation product, a sulfoxide, were the primary excretion products identified in the urine of treated animals (Bartels et al., 2004; Climie et al., 1979; Fisher and Kilgore, 1988a). Approximately 32–36% of acute 5 or 50 mg/kg oral doses of mixed isomers were excreted in male rats in an isomeric ratio of cis- to trans-mercapturate of approximately 4:1 (Bartels et al., 2004). Repeated dosing of rats at 5 mg/kg/day for several weeks appeared to increase the percentage excretion of these metabolites to roughly 43% of the dose (J. Waechter, unpublished data). In contrast, Onkenhout et al. (1986) reported that approximately 45–55% of a range of dosages from 0.05 to 4.5 mg/kg mixed isomers administered via interperitoneal injection to rats were excreted as the mercapturate conjugate at an isomeric ratio of only 1.2:1. Several additional minor metabolites have been observed in the urine of rats and mice (Bartels et al., 2004). These have included the acetate, pyruvate, lactate, and N-acetyl-glutathione conjugates of cis- and trans-1,3-dichloropropene. In addition, Bartels et al. (2004) reported low levels of several novel dimercapturate conjugates of 1,3-dichloropropene in the urine of treated rats and mice. The mercapturic acid conjugate of 1,3-D has also been identified in the urine of human subjects exposed to mixed isomer vapors of 1,3-D under field application or laboratory conditions and has been utilized as a biomonitor of 1,3-D exposure (Kezic et al., 1996; Osterloh and Feldman, 1993; Osterloh et al., 1984; van Welie et al., 1991; Waechter et al., 1992). Consistent with the earlier rat data, Waechter et al. (1992) found that humans excrete approximately 45% and 14% of absorbed cis- and trans-isomers

Hayes’ Handbook of Pesticide Toxicology

of 1,3-dichloropropene, respectively, as the cis-and transmercapturates, a 3.2:1 ratio. The urinary excretion halflife of the mercapturates in humans has been reported to be 5.3 h by van Welie et al. (1991) and 3–5 h by Waechter et al. (1992). Based upon these data, it has been proposed that 1,3dichloropropene is primarily metabolized in rats and mice, and humans, by conjugation with glutathione and by hydrolysis of the 3-position chlorine (Bartels et al., 2004; Dietz et al., 1984; Fisher and Kilgore, 1988a; Hutson et al., 1971; IPCS, 1993; Onkenhout et al., 1986; Osterloh et al., 1984; van Welie et al., 1991; Waechter et al., 1992). The end-product of the latter pathway is CO2. Several other minor metabolites suggestive of oxidative metabolism (epoxidation) have also been reported in the liver of rats administered a very high, lethal, dosage of 1,3dichloropropene (700 mg/kg) via i.p. injection (Schneider et al., 1998). However, a disproportionately lower yield of epoxide was reported at a lower, nonlethal, i.p. dosage and none was detected upon oral dosing (Bartels et al., 2000). Subsequently, Bartels et al. (2004) reported finding approximately 0.6–1.6% of an oral dose of 1,3-dichloropropene excreted in the urine of rats as a 2,3-dimercapturate metabolite. The latter was speculated to possibly represent the product resulting from oxidation of either the parent chemical or of a glutathione conjugate. The toxicity of 1,3dichloropropene thus appears to reflect the balance between inherent chemical reactivity and competing enzymatic activation, and spontaneous and enzymatic detoxification pathways, some or all of which are saturable. A relatively small amount of mixed isomers has been found to bind to macromolecules in the forestomach and, to a lesser extent, in the glandular stomach of rats and mice following acute oral dosing with 1,3-dichloropropene (Dietz et al., 1984). Macromolecular binding correlated with a dose- and time-related depression in the nonprotein sulfhydryl content, presumably of glutathione, as a result of direct or enzymatic conjugation, in stomach and liver tissues of these latter animals. Similar exposure-related decreases in the sulfhydryl content of a number of tissues of rats inhaling 1,3-dichloropropene vapor for1 h have also been reported (Fisher and Kilgore, 1988b). Initial losses in sulfhydryl levels, however, may be offset somewhat by a significant rebound effect, which was observed in livers of rats and lungs of mice repeatedly administered 1,3-dichloro­ propene via oral gavage and inhalation, respectively (W. Stott, unpublished data). Indeed, no DNA adducts were reported by Gollapudi and Stott (2001; reviewed by Stott et al., 2001) in a relatively sensitive 32P-postlabeling assay of these same hepatic and pulmonary tissues and Schneider et al. (1998) reported a sevenfold decrease in formation of epoxide in vitro in the presence of glutathione. Significantly, the net activities of glutathione-S-transferase isozymes to metabolize 1,3-dichloropropene also appear to determine responses in in vitro assays of genotoxicity.

Chapter | 106  1,3-Dichloropropene

In general, target organisms utilized in in vitro assays, especially bacteria, conjugate 1,3-dichloropropene very poorly relative to mammalian tissue extracts, especially when the latter are fortified with physiological levels of glutathione (Creedy et al., 1984; Stott et al., 1992, 2001).

106.3.5  Short-Term Assays 1,3-Dichloropropene has been tested in a wide variety of genotoxicity assays with variable results. A number of early in vitro assays reporting positive responses for 1,3dichloropropene in bacteria (De Lorenzo et al., 1977; Eder et al., 1982; Neudecker et al., 1977; Stolzenberg and Hine, 1980) were confounded by the presence of mutagenic impurities and/or stabilizing agent (e.g., epichlorohydrin) or by the generation of mutagenic oxidation products during gas chromatographic purification procedures (Talcott and King, 1984; Watson et al., 1987). Subsequent studies, using 1,3-dichloropropene purified by passage through a silicic acid column, were weakly positive only in the presence of liver microsomes (100,000  g pellet) from rats induced with polychlorinated biphenyls. Mutagenic activity of impurities was attributed by Eder et al. (2006) to the auto-oxidation in the presence of oxygen of 1,3-dichloropropene to its epoxide followed by chemical degradation, ultimately, to 2-chloroacrolein, a highly mutagenic and genotoxic material. The normal hydrolytic degradation of 1,3-dichloropropene also produces a potential precursor of a different chlorinated aldehyde, 1-chloroallylalcohol (Bartels et al., 2004). Despite this, purified 1-chloroallyl­ alcohol has not been found to be mutagenic in a standard Ames’ Salmonella mutagenicity assay and only weakly mutagenic in a mouse lymphoma assay (B. Gollapudi, unpublished data). Significantly, addition of cytosolic fractions that contain glutathione-S-transferase, and physiological levels of glutathione to in vitro mutagenicity assays, have been found to eliminate mutagenic activity of purified or oxidized samples of 1,3-dichloropropene (Watson et al., 1987; B. Gollapudi, unpublished data). This has also been the case when human cytosol has been utilized in a mutagenicity assay (B. Gollapudi, unpublished data). Metabolically active fractions (S9) obtained from the lungs or kidneys of naïve mice or from the lungs of mice repeatedly exposed to 1,3-dichloropropene vapor also did not metabolize 1,3dichloropropene to a mutagen in Salmonella mutagenicity assays (Gollapudi and Stott, 2001; Stott et al., 1992, 2001). Assays employing mammalian cell lines have also resulted in mixed evidence of genotoxicity. Negative results have been obtained using an epoxidized soybean oil-stabilized mixed isomer formulation of 1,3-dichloropropene in the Chinese hamster ovary (CHO) HGPRT forward mutation and rat hepatocyte unscheduled DNA synthesis assays (Stott et al., 1992, 2001). 1,3-Dichloropropene samples of

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unknown purity or stabilizing agent were negative in chromosomal aberration assays in CHO cells and rat liver cell lines (Dean et al., 1985; Loveday et al., 1989) and negative sister chromatid exchange assays in V79 lung fibroblasts with S9 have also been reported (Loveday et al., 1989). However, similar samples of single or mixed isomeric 1,3-dichloropropene induced unscheduled DNA synthesis in HeLa cells (Schiffman et al., 1983), sister chromatid exchanges in V79 lung fibroblasts (von der Hude et al., 1987), DNA fragmentation and repair in V79 cells and rat and human hepatocytes (Dean et al., 1985; Martelli et al., 1993), chromosomal aberrations in CHL and CHO cells (Loveday et al., 1989), and mutations at the tk locus in L5178Y mouse lymphoma cells (Myhr and Caspary, 1991). In contrast to the results of in vitro genotoxicity assays, in vivo assays of 1,3-dichloropropene have been generally negative. Negative results were obtained in mouse bone marrow micronucleus assays using relatively high oral or i.p. dosages of 1,3-dichloropropene (Gollapudi and Stott, 2001; Shelby et al., 1993). However, Kevekordes et al. (1996) reported that 1,3-dichloropropene induced micro­ nuclei in bone marrow erythrocytes of female, but not male, mice. Inhaled 1,3-dichloropropene vapors did not cause dominant lethal effects in rat germ cells in two separate assays (Gollapudi et al., 1998; Linnett et al., 1988), nor did they cause point mutations in somatic tissues (lung and liver) of Big Blue transgenic mice (Gollapudi and Stott, 2001). In addition, negative results have been obtained for a 1,3-dichloropropene formulation (stabilizer unknown) in several host-mediated bacterial mutagenicity assays in mice (Shirasu et al., 1981; Sudo et al., 1979). A 32P-postlabeling assay of liver tissue from rats dosed orally and of lung tissue of mice exposed via inhalation (target tissues for tumor formation) for the potential formation of DNA adducts was also negative (Gollapudi and Stott, 2001). Positive results, however, have been reported to cause single-strand breaks in the DNA of several tissues and DNA repair in hepatocytes of rats following oral or intraperitoneal dosing (Ghia et al., 1993; Kitchin and Brown, 1994). Finally, when fed to Drosophila at a high concentration, an epoxide-stabilized 1,3-dichloropropene formulation caused an increased incidence of sex-linked recessive lethal mutations but not reciprocal translocations (Valencia et al., 1985).

106.3.6  Chronic Toxicity and Oncogenicity Assays Chronic toxicity and oncogenicity studies of 1,3-dichloropropene via several routes have been conducted. In a recent study using an epoxidized soybean oil-stabilized and weakly in vitro mutagenic formulation of 1,3-dichloropropene, Fischer 344 rats and B6C3F1 mice were administered dosages of 2.5, 12.5, or 25 mg/kg/day and 2.5, 25, or 50 mg/kg/ day, respectively, as a microencapsulated preparation via

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their diet, 7 days/week, for 2 years (Stebbins et al., 2000). In both sexes of rats and mice, body weights were decreased and hyperplasia of the nonglandular stomach mucosa was reported to occur in a dose-related manner. An increased incidence of foci of altered cells was also noted in the livers of treated rats following 24 months dosing. The only tumori­ genic response observed in rats was an increase in the incidence of benign liver tumors in high-dose males and females and intermediate-dose males. No tumorigenic response was reported in either sex of mice. In contrast, an early study involving the administration of an older, highly in vitro mutagenic, mixed-isomer epichlorohydrin-stabilized formulation of 1,3-dichloropropene to rats (25 or 50 mg/kg/day) and mice (50 or 100 mg/kg/ day) via gavage, 3 days/week for up to 2 years resulted in increases in several benign and malignant tumor types in both species (NTP, 1985; Yang et al., 1986). These included forestomach and liver tumors in male rats (25 or 50 mg/kg/day or both), forestomach tumors in female rats (50 mg/kg/day), and forestomach, lung, and urinary bladder tumors in female mice (50 or 100 mg/kg/day or both). The gavage bioassay in male mice was judged to be an “inadequate study of carcinogenicity” due to excessive early mortality of controls. Nontumorigenic responses were limited to hyperplasia of the nonglandular portion of the stomachs of mice and rats and hyperplasia of the urinary bladder epithelium of mice. The chronic toxicity of orally administered epoxidized soybean oil-stabilized formulation of 1,3-dichloropropene has also been evaluated in male and female beagle dogs administered 0.5, 2.5, or 15 mg/kg/day of microencapsulated 1,3-dichloropropene via their diets for 1 year (Stebbins et al., 1999). The primary effect in both sexes of dogs ingesting 15 mg/kg/day 1,3-dichloropropene was a regenerative, hypochromic, microcytic anemia. Histologic changes in bone marrow and spleen consistent with increased hematopoiesis and extramedullary hematopoiesis were consistent with this diagnosis. The latter changes, along with increases in reticulocytes in these animals, confirmed the regenerative nature of this effect. Anemia was observed following 3 months of dosing and remained relatively constant or improved somewhat over the remainder of the dosing period. The only other treatment-related effect observed in the study was a slight inflammation of the tongue of several high-dose males suggestive of irritation by ingestion of 1,3-dichloropropene. Inhalation exposure of rats and mice to 5, 20, or 60 ppm of an epoxidized soybean oil-stabilized formulation of 1,3dichloropropene 6 h/day, 5 days/week, for 2 years resulted in nontumorigenic lesions of the nasal mucosa in both sexes of rats and mice exposed to 60 ppm and female mice exposed to 20 ppm, the urinary bladder epithelium of both sexes of mice exposed to 60 ppm and the forestomach of male mice exposed to 60 ppm (Lomax et al., 1988). Slight changes in the morphology of renal and hepatic tissues of male and

Hayes’ Handbook of Pesticide Toxicology

female mice exposed to 60 ppm, respectively, indicative of decreased lipid and glycogen content, respectively, were also observed. An increased incidence of benign lung tumors in high-exposure group male mice was the only tumorigenic response observed (44% vs 18% in controls; historical incidence of adenomas, 7–32%). cis-1,3-Dichloropropene was negative in a mouse skin initiation-promotion bio­assay when tested with phorbol myristate as a promoter and was not carcinogenic following repeated dermal application of 122 mg to the backs of Ha:ICR Swiss mice three times/week for up to 85 weeks (Van Duuren et al., 1979). An increase in local fibrosarcomas (six of 30 mice vs. 0 of 30 controls) was reported in mice following repeated subcutaneous injections of 3 mg/animal/week (1/week) for up to 83 weeks.

106.4  Dose response The toxicity of 1,3-dichloropropene in a number of laboratory animals displays both a clear dose response and clearly defined no-observed-effect levels (NOELs). This is true for both acute and repeated dose (subacute, subchronic, and chronic) nonneoplastic treatment-related effects as well as neoplastic effects in rodents upon chronic exposure (see above discussion of individual studies). A number of toxi­ city studies with their lowest effect levels (LEL), NOELs, and target tissues/effect at the LELs are summarized in Table 106.1. It can be seen that the most sensitive target tissues/effects (i.e., those observed at LELs) are consistent between studies of differing durations for a given species of test animal. Increasing the duration of the dosing period from 2–13 weeks to 2 years did not appear to significantly change the potential toxicity of 1,3-dichloropropene relative to nonneoplastic pathological effects. Changes in NOELs or no-observed-adverse-effect levels (NOAELS) obtained in studies in which 1,3-dichloropropene was ingested or inhaled were generally within less than half a log unit of each other. An exception to this was the series of oral dietary toxicity studies in mice in which body weight depression was the major treatment-related change noted. In this case, more significant duration of treatmentdependent decreases in NOELs occurred. The most sensitive treatment-related effects observed in animals ingesting or inhaling 1,3-dichloropropene were quite similar between studies for a given species and method of administration. Affected tissues often represented portal-of-entry tissues, for example gastric mucosa for ingested and nasal mucosa for inhaled 1,3-dichloropropene, consistent with the irritant nature of this chemical. Exceptions to this were the occurrence of hyperplasia of the transitional epithelium of the urinary bladders of both sexes of mice inhaling vapors or dosed orally via gavage, and anemia in dogs ingesting 1,3-dichloropropene. In most bioassays, tumorigenic responses, when present, have

Sex

Study duration

LELa

NOELb

Target tissue at LEL

Reference

Males Females Males Females Males Females

2 Weeks

25 mkdc 50 mkd 15 mkd 15 mkd 12.5 mkd 12.5 mkd

10 mkd 25 mkd 5* mkd 5 mkd 2.5 mkd 2.5 mkd

Body wt., gastric mucosa Body wt., gastric mucosa Body wt., gastric mucosa Body wt., gastric mucosa Body wt., gastric mucosa, liver ADd Body wt., gastric mucosa

Haut et al. (1992b)

Males Females Males and females Males and females

2 Weeks

100 mkd 175 mkd 50 mkd 25 mkd

50 mkd 100 mkd 15* mkd 2.5 mkd

Body wt. Body wt. Body wt. Body wt.

Haut et al. (1992a)

Males Females Males and females

13 Weeks

15 mkd 16 mkd 15 mkd

5* mkd 5* mkd 2.5 mkd

Anemia Body wt., anemia Anemia

Stebbins et al. (1999)

Rats

Males Females

2 Years

25 mkd 25 mkd

NDe ND

Gastric mucosa, liver AD Gastric mucosa, stomach AD

NTP (1985)

Mice

Females

2 Years

50 mkd

ND

Gastric and u. bladder (CAf) mucosa, lung AD and CA

NTP (1985)

Males and females Males Females Males and females Males and females

4 Weeks 13 Weeks

30 ppm 30 ppm 10 ppm 10 ppm 20 ppm

ND Body wt., nasal mucosa Nasal mucosa Body wt., (nasal mucosa) Body wt., nasal

Coate (1979b) Coate (1979a); Stott et al. (1988)

13 Weeks 2 Years

ND 90 ppm 30 ppm 30 ppm (90 ppm)g 60 ppm

Males and females Males and females Males and females Males and females

4 Weeks 13 Weeks 13 Weeks 2 Years

ND 90 ppm 30 ppm 20 ppm

30 ppm 30 ppm 10 ppm 5 ppm

ND Body wt., nasal mucosa (females) Body wt., nasal & urinary bladder mucosa Nasal mucosa

Coate (1979b) Coate (1979a); Stott et al. (1988) Stott et al. (1988) Lomax et al. (1988)

Route species Oral (diet) Rats

Mice

Dogs

13 Weeks 2 Years

13 Weeks 2 Years

1 Year

Haut et al. (1996) Stebbins et al. (2000) Stebbins et al. (2000)

Chapter | 106  1,3-Dichloropropene

Table 106.1  Summary of Lowest Effective Levels, No-Observed-Effect Levels, and Most Sensitive Treatment-Related Effects for Toxicity Studies of Mixed and cis-Isomer 1,3-Dichloropropene Formulations

Haut et al. (1996) Haut et al. (1996)

Stebbins et al. (1999)

Oral (gavage)

Inhalation Rats

Mice

Stott et al. (1988) Lomax et al. (1988)

*

Signifies no-observed-adverse-effect level. Lowest effect level. No-observed-effect level. c mkd  mg/kg/day. d AD  adenoma (benign tumors). e ND  not determined. f CA  carcinomas (malignant tumors). g Exposure level in parenthesis reflects exposure level at which effects on nasal mucosa were observed. a

b

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involved portal-of-entry tissues. Ingestion or inhalation of relatively recent formulations of 1,3-dichloropropene resulted in tumors in the livers, often regarded as a portalof-entry tissue of the enteric tract, of rats and the lungs of male mice, respectively. However, chronic gavage of an older and highly mutagenic, epoxide-stabilized formulation of 1,3-dichloropropene, while causing forestomach and liver tumors in rats and mice, also caused urinary bladder and lung tumors in mice. The effect of 1,3-dichloropropene upon portal-of-entry tissues is consistent with a direct toxicity of the molecule and its removal by saturable metabolic pathways. As noted, a major pathway for 1,3-dichloropropene metabolism is via glutathione-S-transferase-dependent conjugation with glutathione. As reviewed by Watson et al. (1987) and in the IPCS (1993) review, this metabolism provides for practical thresholds in the dose response of toxicity, and even mutagenicity, of 1,3-dichloropropene in test organisms. The saturation of this pathway may result in a nonlinear elevation in concentrations of 1,3-dichloropropene in cells of an in vitro mutagenicity or clastogenicity assay or tissues of an exposed animal, especially in portal-of-entry tissues, and subsequent toxicological consequences.

106.5  Toxicology in humans 106.5.1  Experimental Exposure Seven out of 10 volunteers detected 1,3-dichloropropene at an air concentration of 3 ppm; some reported fatigue of the sense of smell after a few minutes. The same proportion of volunteers detected 1 ppm, but the odor was noticeably fainter (Torkelson and Oyen, 1977). In a population of 22 persons, the concentration at which odor was detected was 4.4  3.1 ppm (mean  S.D.) (Rick and McCarty, 1987).

106.5.2  Accidental Poisoning Forty-six people were treated for exposure to 1,3dichloropropene fumes following a traffic accident in 1975 involving spillage of 4500 l of a formulated product (Flessel et al., 1978). Twenty-four of these, three of whom had lost consciousness, were hospitalized overnight with symptoms including headache, irritation of mucous membranes, and chest discomfort. All patients took showers and were given intravenous fluids and three received oxygen and corticosteroids because of chest pain and cough. Eleven of 41 persons tested had slightly higher than average serum SGOT and/or SGPT values that reverted to normal within 48–72 h, except for five who still had slightly higher than average SGOT values. Follow-up interviews with patients 1–2 weeks later revealed symptoms including headache, abdominal, and chest discomfort, and malaise. One was diagnosed as having had pneumonia. Symptoms

Hayes’ Handbook of Pesticide Toxicology

were reported more frequently in those most heavily exposed to the fumes. Patient interviews conducted approximately 2 years after the accident revealed complaints of headache, chest pain or discomfort, and “personality changes” (fatigue, irritability, difficulty in concentrating, or decreased libido). Two had undergone cardiac catheterizations but their arteriograms were normal. There was no correlation of these long-persisting symptoms with intensity of exposure. Two fatalities involving 1,3-dichloropropene have been confirmed. Accidental ingestion of D-D (admixture with 1,2dichloropropane) resulted in abdominal pain and vomiting, muscular twitching, pulmonary edema, and death (Gosselin et al., 1976). Accidental ingestion of TELONE II by a farm worker in Spain resulted in abdominal pain and vomiting, adult respiratory distress syndrome, hematologic changes, hepatorenal impairment, muscular twitching followed by coma, and death (Hernandez et al., 1994). A possible association between overexposure to 1,3-dichloropropene and the development of hematologic malignancy has also been suggested by Markovitz and Crosby (1984). The latter was based upon the development of histiocytic lymphoma in a farm worker and two firemen accidentally exposed (acute) to high levels TELONE II and D-D, respectively; however, it was subsequently established that the farm worker had leukemia prior to the incident (public records, State of California, Court of Appeal, Case No. 28344).

106.5.3  Use Experience 1,3-Dichloropropene causes edema, redness, and necrosis of the skin (Torkelson and Oyen, 1977) and in one documented case was believed to have caused a contact hypersensitivity in a repeatedly exposed farmer (Nater and Gooskens, 1976). A fertility study of 64 employees engaged in the production of chlorinated 3-carbon compounds, including 1,3-dichloropropene, revealed no effects upon hormone levels (LH, FSH, testosterone), sperm count, sperm motility, and percent normal and abnormal sperm regardless of duration or magnitude of exposure (Venable et al., 1980). Several studies have also been undertaken to evaluate potential biological effects of 1,3-dichloropropene vapors in fumigation workers. In a California study, urinary parameters were measured over time in the same workers following single or repeated occupational exposure(s) to a range of 0.3–9.4 mg/m3 mixed isomers of 1,3-dichloropropene during soil fumigation operations (Osterloh and Feldman, 1993). Exposure was reportedly associated with elevated urinary excretion of N-acetylglucosamidase (NAG) and retinol-binding protein (RBP). No changes were observed in albumin (ALB) excretion. These findings were interpreted to suggest a “subclinical” renal toxicity. In a Dutch study, several urinary and serum parameters were measured in the same workers occupationally exposed to

Chapter | 106  1,3-Dichloropropene

1.9–18.9 mg/m3 cis-1,3-dichloropropene products once before and once after the tulip bulb field fumigation season (about a 3-month period) (Brouwer et al., 1991). A number of slight changes in several parameters were reported after relative to before the “season”; excretion of urinary NAG and ALB, decreased serum creatinine (CREAT-S), and decreased total bilirubin (TBILI) levels in combination with increased serum -glutamyltranspeptidase (GGT). No differences were reported in serum 2-microglobulin (2M), alanine aminopeptidase (AAP), -galactosidase, alkaline phosphatase (ALP), aspartate aminotransferase (AST), alanine aminotransferase (ALT), or lactate dehydrogenase. These data were interpreted by Brouwer et al. (1991) to reflect a slight degree of liver and kidney toxi­ city. Both of these studies have been strongly criticized for perceived study design and data interpretation flaws (Stott et al., 1990; van Sittert et al., 1991). A subsequent comprehensive study in Dutch potato field fumigation workers conducted over the whole of the fumigation season failed to observe treatment-related toxicity (Verplanke et al., 2000). Workers were exposed to a range of 0.1–9.5 mg/m3 cis-1,3-dichlororporpene and a number of parameters were measured before, during, and following the fumigation season. Unlike previous studies, this study employed a matched control group of individuals. Parameters measured included urinary AAP, NAG, RBP, and ALB, and serum 2M, CREAT-S, ALT, AST, GGT, ALP, and TBILI. The only change observed was a slightly lower urinary ratio of 6--hydroxycortisol to free cortisol ratio, which was not considered to be related to 1,3-dichloropropene exposure. It was concluded that no adverse effects on liver or kidney function were suggested by the data. Additional reviews of the human toxicity data for 1,3dichloropropene have been published by Yang (1986) and IPCS (1993).

106.6  Summary risk characterization 1,3-Dichloropropene has found use for over 45 years and remains one of the few remaining compounds available to agriculture for fumigating soils to eliminate parasitic nematodes. This compound has been extensively evaluated in a number of test organisms for acute, subchronic, and chronic toxicity, reproductive and developmental toxicity, carcinogenicity, and genotoxicity. Its metabolism in animals, including humans, has also been extensively studied and is relatively well understood. 1,3-Dichloropropene is moderately to highly acutely toxic to animals. It is irritating to skin and if occluded can cause a chemical burn and death in rabbits; however, its relatively high vapor pressure results in much lower toxicity if left on skin unoccluded. Orally administered 1,3-dichloropropene, either neat or

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in an aqueous vehicle, may be lethal to rodents at roughly 100 mg/kg or greater. Two human fatalities from accidental imbibition of 1,3-dichloropropene formulations, a relatively purified 1,3-dichloropropene and an admixture with 1,2-dichloropropane, have been reported. Toxic effects observed in humans exposed to high levels of vapors or having extended dermal contact with the liquid appear to reflect the irritant nature of 1,3-dichloropropene. In animal models, the results of subchronic and chronic toxicity studies of this chemical reflect both its irritant properties, as evidenced by effects on portal-of-entry tissues, and potential toxicity of a metabolite(s), as evidenced by effects upon distal tissues (e.g., urinary bladder mucosa). Despite this, studies have demonstrated a lack of reproductive or developmental effects, even at toxic dosages. 1,3-Dichloropropene is rapidly and extensively metabolized upon absorption by animals, including humans. Elimination from the blood of rats occurs in a biphasic manner, with half-lives for both isomers in the prominent -phase of approximately 4–7 min and the -phase of approximately 25–45 min. No appreciable excretion of parent chemical occurs and metabolites are primarily eliminated via the urine as products of a glutathione conjugation metabolic pathway or via exhalation of CO2, product of a hydrolytic pathway. Evidence of the former pathway has been dose-related decreases in tissue glutathione levels of rats and mice administered 1,3-dichloropropene via oral or inhalation routes. The major urinary metabolite, the mercapturate conjugate of 1,3-dichloropropene, has represented a useful biomarker by which to estimate the exposure of workers to this molecule during soil fumigation operations. Genotoxicity tests of 1,3-dichloropropene have often provided contradictory results. Many short-term assays of genotoxicity have been confounded by the presence of a known mutagen, epichlorohydrin, in the formulated material tested, which was historically added as a stabilizing agent. The potential of 1,3-dichloropropene to undergo auto-oxidation to generate a mutagenic epoxide has further complicated interpretation of the in vitro genotoxicity data. Epoxide may be formed upon prolonged exposure to oxygen or during gas-chromatographic “purification” procedures carried out prior to testing. In vivo assays of mutagenic or clastogenic activity, with their intact complement of metabolizing enzymes, have almost uniformly been negative, especially at nontoxic dosages or at dosages that do not deplete tissue glutathione levels. It has been proposed that the genotoxic potential of 1,3-dichloropropene is directly related to the extensive depletion of glutathione in target organisms and tissues (IPCS, 1993). Based upon a weight-of-the-evidence analysis inclusive of in vivo assay data, 1,3-dichloropropene lacks significant genotoxic activity. Bioassay data have provided an equally complicated assessment of the potential of 1,3-dichloropropene to cause

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tumors in animals. Inhalation of vapor, the primary route of occupational exposure to this chemical, has been shown to cause an increase in the incidence of benign lung tumors in male mice. Ingestion of this chemical via the diet as a stabilized microencapsulated product has been shown to cause a low incidence of benign liver tumors in rats. These results contrast with those of a previous oral oncogenicity study conducted by repeated bolus dosing (gavage) of an older, epichlorohydrin stabilized, and highly mutagenic formulation of 1,3-dichloropropene, which resulted in numerous benign and malignant tumors in both rats and mice.

References Bartels, M. J., Brzak, K. A., Mendrala, A. L., and Stott, W. T. (2000). Mechanistic aspects of the metabolism of 1,3-dichloropropene in rats and mice. Chem. Res. Toxicol. 13, 1096–1102. Bartels, M. J., Hansen, S. C., Thornton, C. M., Brzak, K. A., Mendrala, A. L., Dietz, F. K., and Kastl, P. E. (2004). Pharmacokinetics and metabolism of 14C-1,3-dichloropropene in the Fischer 344 rat and B6C3F1 mouse. Xenobiotica 34, 193–213. Breslin, W. J., Kirk, H. D., Streeter, C. M., Quast, J. F., and Szabo, J. R. (1989). 1,3-Dichloropropene: Two-generation inhalation reproduction study in Fischer 344 rats. Fund. Appl. Pharmacol. 12, 129–143. Brouwer, E. J., Evelo, C. T. A., Verplanke, A. J. W., van Welie, R. T. H., and de Wolff, F. A. (1991). Biological effect monitoring of occupational exposure to 1,3-dichloropropene: Effects on liver and renal function and on glutathione conjugation. Br. J. Ind. Med. 48, 167–172. Bruckner, J. V., MacKenzie, W. F., Ramanathan, R., Muralidhara, S., Kim, H. J., and Dallas, C. E. (1989). Oral toxicity of 1,2-dichloropropane: Acute, short-term and long-term studies in rats. Fund. Appl. Toxicol. 12, 713–730. Climie, I., Hutson, D., Morrison, B., and Stoydin, G. (1979). Glutathione conjugation in the detoxication of (Z)-1,3-dichloropene (a component of the nematocide D-D) in the rat. Xenobiotica 9, 149–156. Coate, W. B. (1979a). 90-Day Inhalation Study in Rats and Mice: TELONE II. Report of The Dow Chemical Company. Coate, W. B. (1979b). Subacute Inhalation Study in Rats and Mice: TELONE II. Report of The Dow Chemical Company. Creedy, C., Brooks, T., Dean, B., Hutson, D., and Wright, A. (1984). The protective action of glutathione on the microbial mutagenicity of the Z- and E-isomers of 1,3-dichloropropene. Chem. Biol. Interact. 50, 39–48. Dean, B. J., Brooks, T. M., Hodson-Walker, G., and Hutson, D. H. (1985). Genetic toxicology testing of 41 industrial chemicals. Mut. Res. 153, 57–77. De Lorenzo, F., Degl’Innocenti, S., Ruocco, A., Silengo, L., and Cotese, R. (1977). Mutagenicity of pesticides containing 1,3-dichloropropene. Cancer Res. 37, 1915–1917. Dietz, F., Dittenber, D., Kirk, H., Ramsey, J. (1984). Non-protein sulfhydryl content and macromolecular binding in rats and mice following administration of 1,3-dichloropropene. Toxicologist 4, Abst. No. 586. Eder, E., Neudecker, T., Lutz, D., and Henschler, D. (1982). Correlation of alkylating and mutagenic activities of allyl and allylic compounds: standard alkylation test vs. kinetic investigation. Chem. Biol. Interact. 38, 303–315. Eder, E., Espinosa-Gonzalez, J., Mayer, A., Reichenberger, K., and Boerth, D. (2006). Autooxidative activation of the nematocide 1,3-dichloropropene to highly genotoxic and mutagenic derivatives: Consideration

Hayes’ Handbook of Pesticide Toxicology

of genotoxic/carcinogenic mechanisms. Chem. Res. Toxicol. 19, 952–959. Fisher, G. D., and Kilgore, W. W. (1988a). Mercapturic acid excretion by rats following inhalation exposure to 1,3-dichloropropene. Fundam. Appl. Toxicol. 11, 300–307. Fisher, G. D., and Kilgore, W. W. (1988b). Tissue levels of glutathione following acute inhalation of 1,3-dichloropropene. J. Toxicol. Environ. Health 23, 171–182. Flessel, P., Goldsmith, J., Kahn, E., Wesolwski, J., Maddy, K., and Peoples, S. (1978). Acute and possible long-term effects of 1,3dichloropropene – California. Morb. Mortal. Wkly. Rep. 27, 50–55. Gardner, J. R. (1989). cis-1,3-Dichloropropene: Acute Oral and Dermal Toxicity, Skin and Eye Irritancy and Skin Sensitisation Potential. Report of Sittingbourne Research Center. Ghia, M., Robbiano, L., Allavena, A., Martelli, A., and Brambilla, G. (1993). Genotoxic activity of 1,3-dichloropropene in a battery of in vivo short-term tests. Toxicol. Appl. Pharmacol. 120, 120–125. Gollapudi, B. B., Stott, W. T. (2001). Cis/trans 1,3-dichloropropene is not genotoxic in vivo in tumor target and non-target tissues, Toxicologist 60, Abstr. No. 729. Gollapudi, B. B., Cieszlak, F. S., Day, S. J., and Carney, E. W. (1998). Dominant lethal test with rats exposed to 1,3-dichloropropene. Environ. Mol. Mutagen. 32, 351–359. Gosselin, R., Hodge, H., Smith, R., and Gleason, M. (1976). “Clinical Toxicology of Commercial Products,” 4th ed, pp. 119–121 WilkinsWilliams, Baltimore. Hanley, T. R., John-Greene, J. A., Young, J. T., Calhoun, L. L., and Rao, K. S. (1987). Evaluation of the effects of inhalation exposure to 1,3dichloropropene on fetal development in rats and rabbits. Fundam. Appl. Toxicol. 8, 562–570. Haut, K. T., Stebbins, K. E., Kropscott, B. E., Stott, W. T. (1992a). TELONE®II Soil Fumigant: Palatability and Two-Week Dietary Probe Studies in B6C3F1 Mice. Report of The Dow Chemical Company, Midland, MI. Haut, K. T., Stebbins, K. E., Kropscott, B. E., Stott, W. T. (1992b). TELONE®II Soil Fumigant: Palatability and Two-Week Dietary Probe Studies in Fischer 344 Rats. Report of The Dow Chemical Company, Midland, MI. Haut, K. T., Johnson, K. A., Shabrang, S. N., and Stott, W. T. (1996). Subchronic toxicity of ingested 1,3-dichloropropene in rats and mice. Fundam. Appl. Toxicol. 32, 224–232. Hernandez, A. F., Martin-Rubi, J. C., Ballesteros, J. L., Oliver, M., Pla, A., and Villanueva, E. (1994). Clinical and pathological findings in fatal 1,3-dichloropropene intoxication. Hum. Exp. Toxicol. 13, 303–306. Hine, C. H., Anderson, H. H., Moon, H. D., Kodama, J. K., Morse, M., and Jacobson, N. W. (1953). Toxicology and safe handling of CBPSS (technical 1-chloro-3-bromopeopene-1). Arch. Ind. Hyg. Occup. Med. 7, 118–136. Hutson, D. H., Moss, J. A., and Pickering, B. A. (1971). Components of the soil fumigant D-D* and their metabolites in the rat. Food Cosmet. Toxicol. 9, 677–680. IPCS (International Programme on Chemical Safety) (1993). 1,3Dichloropropene, 1,2-Dichloropropane and Mixtures. Environmental Health Criteria No. 146, World Health Organization, Geneva. Jones, J. R. (1988a). 1,3-Dichloropropene cis-Isomer: Acute Dermal Toxicity Test in the Rat. Report of Safepharm Laboratories Limited, Derby, UK. Jones, J. R. (1988b). 1,3-Dichloropropene cis-Isomer: Acute Oral Toxicity Test in the Rat. Report of Safepharm Laboratories Limited, Derby, UK.

Chapter | 106  1,3-Dichloropropene

Jones, J. R. (1988c). 1,3-Dichloropropene cis-Isomer: Modified NineInduction Buehler Contact Sensitisation Study in the Guinea Pig. Report of Safepharm Laboratories Limited, Derby, UK. Jones, J. R., and Collier, T. A. (1986a). TELONE II: Acute Dermal Toxicity Test in the Rat. Report of The Dow Chemical Company. Jones, J. R., and Collier, T. A. (1986b). TELONE II: Acute Oral Toxicity in the Rat. Report of The Dow Chemical Company. Kevekordes, S., Gebel, T., Pav, K., Edenharder, R., and Dunkelberg, H. (1996). Genotoxicity of selected pesticides in the mouse bone marrow micronucleus test and in the sister-chromatid exchange test with human lymphocytes in vitro. Toxicol. Lett. 89, 35–42. Kezic, S., Monster, A. C., Verplanke, J. W., and de Wolff, F. A. (1996). Dermal absorption of cis-1,3-dichloropropene vapour: human experimental exposure. Hum. Exp. Toxicol. 15, 396–399. Kitchin, K. T., and Brown, J. L. (1994). Dose-response relationship for rat liver DNA damage caused by 49 rodent carcinogens. Toxicology 88, 31–49. Linnett, S. L., Clark, D. G., Blair, D., and Cassidy, S. L. (1988). Effects of subchronic inhalation of D-D (1,3-dichloropropene/1,2-dichloropropane) on reproduction in male and female rats. Fundam. Appl. Toxicol. 10, 214–223. Lomax, L. G., Stott, W. T., Johnson, K. A., Calhoun, L. L., Yano, B. L., and Quast, J. F. (1988). The chronic toxicity and oncogenicity of inhaled technical grade 1,3-dichloropropene in rats and mice. Fundam. Appl. Toxicol. 12, 418–431. Loveday, K. S., Lugo, M. H., Resnick, M. A., Anderson, B. E., and Zeiger, E. (1989). Chromosome aberration and sister chromatid exchange tests in Chinese hamster ovary cells in vitro: II. Results with 20 chemicals. Environ. Mol. Mutagen. 13, 60–94. Markovitz, A., and Crosby, W. H. (1984). Chemical carcinogenesis. A soil fumigant, 1,3-dichloropropene, as possible cause of hematologic malignancies. Arch. Intern. Med. 144, 1409–1411. Martelli, A., Allavena, A., Ghia, M., Robbiano, L., and Brambilla, G. (1993). Cytotoxic and genotoxic activity of 1,3-dichlororporpene in cultured mammalian cells. Toxicol. Appl. Pharmacol. 120, 114–119. Myhr, B. C., and Caspary, W. J. (1991). Chemical mutagenesis at the thymidine kinase locus in L5178Y mouse lymphoma cells: Results for 31 coded compounds in the National Toxicology Program. Environ. Mol. Mutagen. 18, 51–83. Nater, J. P., and Gooskens, V. H. J. (1976). Occupational dermatosis due to a soil fumigant. Contact Dermat. 2, 227–229. National Toxicology Program (NTP). (1985). Toxicology and Carcinogenesis Studies of TELONE II® in F344/N Rats and B6C3F1 Mice (Gavage Studies). NTP Tech. Rep. 269. Government Printing Office, Washington, DC. Neudecker, T., Stefani, A., and Henschler, D. (1977). In vitro mutagenicity of soil nematocide 1,3-dichloropropene. Experientia 33, 1084–1085. Nitschke, K. D., and Lomax, L. G. (1990). Cis-1,3-Dichloropropene: 2Week Vapor Inhalation Toxicity Study in Fischer 344 Rats. Report of The Dow Chemical Company, Midland, MI. Nitschke, K. D., Crissman, J. W., and Schuetz, D. J. (1990a). Cis-1,3Dichloropropene: Acute Inhalation Toxicity Study with Fischer 344 Rats. Report of The Dow Chemical Company, Midland, MI. Nitschke, K. D., Lomax, L. G., and Sanderson, T. G. (1990b). Cis-1,3Dichloropropene: 13-Week Vapor Inhalation Toxicity Study in Fischer 344 Rats. Report of The Dow Chemical Company, Midland, MI. Onkenhout, W., Mulder, P. P. J., Boogaard, P. J., Buijs, W., and Vermeulen, N. P. E. (1986). Identification and quantitative determination of mercapturic acids formed from Z- and E-1,3-dichloropropene by the rat, using gas chromatography with three different detection techniques. Arch. Toxicol. 59, 235–241.

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Osterloh, J. D., and Feldman, B. J. (1993). Urinary protein markers in pesticide applicators during a chlorinated hydrocarbon exposure. Environ. Res. 63, 171–181. Osterloh, J., Letz, G., Pond, S., and Becker, C. (1983). An assessment of the potential testicular toxicity of 10 pesticides using the mousesperm morphology assay. Mutat. Res. 116, 407–415. Osterloh, J. D., Cohen, B. S., Popendorf, W., and Pond, S. M. (1984). Urinary excretion of the N-acetyl cysteine conjugate of cis-1,3-dichloropropene by exposed individuals. Arch. Environ. Health 39, 271–275. Parker, C., Coate, W., and Voelker, R. (1982). Subchronic inhalation toxicity of 1,3-dichloropropene/1,2-dichloropropane (D-D®) in mice and rats. J. Toxicol. Environ. Health 9, 899–910. Rick, D. L., and McCarty, L. P. (1987). The Determination of the Odor Threshold of Vapors and Gases. Report of The Dow Chemical Company. Schiffman, D., Eder, E., Neudecker, T., and Henschler, D. (1983). Induction of unscheduled DNA synthesis in HeLa cells by allylic compounds. Cancer Lett. 20, 263–269. Schneider, M., Quistad, G. B., and Casida, J. E. (1998). 1,3-Dichloropropene epoxides: Intermediates in bioactivation of the promutagen 1,3-dichloropropene. Chem. Res. Toxicol. 11, 1137–1144. Shelby, M. D., Erexson, G. L., Hook, G. J., and Tice, R. R. (1993). Evaluation of a three-exposure mouse bone marrow micronucleus protocol: Results with 49 chemicals. Environ. Mol. Mutagen. 21, 160–179. Shirasu, Y., Moriya, M., Tequka, H., Teramoto, S., Ohata, T., and Inoue, T. (1981). Mutagenicity screening studies on pesticides. In “Environmental Mutagens and Carcinogens: Proceedings of the Third International Conference on Environmental Mutagens”. Tokyo, Mishima, and Kyoto, September 21–27, pp. 331–335. Stebbins, K. E., Quast, J. F., Haut, K. T., and Stott, W. T. (1999). Subchronic and chronic toxicity of ingested 1,3-dichloropropene in beagle dogs. Regul. Toxicol. Pharmacol. 30, 233–243. Stebbins, K. E., Johnson, K. A., Jeffries, T. K., Redmond, J. M., Haut, K. T., Shabrang, S. N., and Stott, W. T. (2000). Chronic toxicity and oncogenicity studies of ingested 1,3-dichloropropene in rats and mice. Regul. Toxicol. Pharmacol. 32, 1–13. Stolzenberg, S., and Hine, C. (1980). Mutagenicity of 2- and 3-carbon halogenated compounds in the Salmonella/mammalian-microsome test. Environ. Mutagen. 2, 59–66. Stott, W. T., and Kastl, P. L. (1986). Inhalation pharmacokinetics of technical grade 1,3-dichloropropene in rats. Toxicol. Appl. Pharmacol. 85, 332–341. Stott, W., Young, J., Calhoun, L., and Battjes, J. (1988). Subchronic toxi­ city of inhaled technical grade 1,3-dichloropropene in rats and mice. Fundam. Appl. Toxicol. 11, 207–220. Stott, W. T., Waechter, J. M., and Quast, J. T. (1990). Letter to the editor. Arch. Environ. Health 45, 250–253. Stott, W. T., Mendrala, A. M., Redmond, J. M., Nwosu, A. F., and Lomax, L. G. (1992). Mechanism of 1,3-dichloropropene (1,3-D) induced toxicity in urinary bladder epithelium of mice. Toxicologist 12, Abstr. No. 415. Stott, W. T., Gilbert, J. R., McGuirk, R. J., Brzak, K. A., Alexander, L. M., Dryzga, M. D., Mendrala, A. L., and Bartels, M. J. (1998). Bioavailability and pharmacokinetics of microencapsulated 1,3dichloropropene in rats. Toxicol. Sci. 41, 21–28. Stott, W. T., Gollapudi, B. B., and Rao, K. S. (2001). A review of the mammalian toxicity of 1,3-dichloropropene. Rev. Environ. Contam. Toxicol. 168, 1–42. Sudo, S., Kimura, Y., Yamamoto, K., and Ichihara, S. (1979). The Mutagenicity Test on 1,3-Dichloropropene in Bacteria Test System. Report of The Dow Chemical Company, Midland, MI.

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Talcott, R., and King, J. (1984). Mutagenic impurities in 1,3-dichloropropene preparations. J. Natl. Cancer Inst. 72, 1113–1116. Til, H. P., Spanjers, M. T., Feron, V. J., Reuzel, P. J. C. (1973). Subchronic (90-Day) Toxicity Study with TELONE* in Albino Rats. Report of The Dow Chemical Company, Horgen, Switzerland. Torkelson, T. R. (1994). Halogenated aliphatic hydrocarbons containing chlorine, bromine, and iodine. In “Patty’s Industrial Hygeine and Toxicology” (G. D. Clayton and F. E. Clayton, eds.), 4th ed., pp. 4007–4251. Wiley, New York. Torkelson, R., and Oyen, F. (1977). The toxicity of 1,3-dichloropropene as determined by repeated exposure of laboratory animals. Am. Ind. Hyg. Assoc. J. 38, 217–223. Valencia, R., Mason, J. M., Woodruff, R. C., and Zimmering, S. (1985). Chemical mutagenesis testing in Drosophila. III. Results of 48 coded compounds tested for the national toxicology program. Environ. Mutagen. 7, 325–348. Van Duuren, B. L., Goldschmidt, B. M., Loewengart, G., Smith, A. C., Melchionne, S., Seidman, I., and Roth, D. (1979). Carcinogenicity of halogenated olefinic and aliphatic hydrocarbons in mice. J. Natl. Cancer Inst. 63, 1433–1439. van Sittert, N. J., Veenstra, G. E., Dumas, E. P., and Tordoir, E. F. (1991). Letter to the editor. Br. J. Med. 48, 646–648. van Welie, R. T. H., van Duyn, P., Brouwer, D. H., van Hemmen, J. J., Brouwer, E. J., and Vermeulen, N. P. E. (1991). Inhalation exposure to 1,3-dichloropropene in the Dutch flower-bulb culture. Part II. Biological

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monitoring by measurement of urinary excretion of two mercapturic acid metabolites. Arch. Environ. Contam. Toxicol. 20, 6–12. Venable, J. R., McClimans, C. D., Flake, R. E., and Dimick, D. B. (1980). A fertility study of male employees engaged in the manufacture of glycerine. J. Occup. Med. 22, 87–91. Verplanke, A. J. W., Bloemen, L. J., Brouwer, E. J., Van Sittert, N. J., Boogaard, P. J., Herber, R. F. M., and DeWolff, F. A. (2000). Occupational exposure to cis-1,3-dichloropropene: Biological effect monitoring of kidney and liver. Occup. Environ. Med. 57, 745–751. von der Hude, W., Scheutwinkel, M., Gramlich, U., Fibler, B., and Basler, A. (1987). Genotoxicity of three-carbon compounds evaluated in the SCE test in vitro. Environ. Mutagen. 9, 401–410. Watson, W. P., Brooks, T. M., Huckle, K. R., Hutson, D. H., Land, K. L., Smith, R. J., and Wright, A. S. (1987). Microbial mutagenicity studies with (Z)-1,3-dichloropropene. Chem. Biol. Interact. 61, 17–30. Waechter, J. M., Brzak, K. A., McCarty, L. P., LaPack, M. A., and Brownson, P. J. (1992). Cis/trans 1,3-dichloropropene (1,3-dichloropropene): Inhalation pharmacokinetics and metabolism in human volunteers. Toxicologist 13, Abstr. No. 1090. Yang, R. S. H. (1986). 1,3-dichloropropene. Residue Rev. 97, 19–35. Yang, R. S. H., Huff, J. E., Boorman, G. A., Haseman, J. K., Kornreich, M., and Stookey, J. L. (1986). Chronic toxicology and carcinogenesis studies of TELONE II by gavage in Fischer 344 rats and B6C3F1 mice. J. Toxicol. Environ. Health 18, 377–392.

Chapter 107

Metam-Sodium Linda L. Carlock1 and Timothy A. Dotson2 1 2

Toxicology and Regulatory Consulting UCB Chemicals Corporation

107.1  Introduction Metam-sodium (C2H4NNaS2, CAS no. 137-42-8), also known as metham sodium, sodium metam, sodium-N-methyldithiocarbamate, methylcarbamodithioic acid sodium salt, methyldithiocarbamic acid sodium salt, carbam, and SMDC, is a white crystalline powder in the pure form but is normally found as a clear yellow liquid with a strong sulfurlike odor (Merck, 1989). Metam-sodium is prepared from methylamine, carbon disul-fide, and sodium hydroxide in an aqueous solution. Metam-sodium has a molecular weight of 129.18. Metam-sodium is stable in its dry, crystalline state, and in concentrated aqueous solution. In solution, metam-sodium has a vapor pressure of 21 mg Hg at 25°C (U.S. EPA, 1994a). Metam-sodium is very stable at a pH greater than 8.8, but at pH 7 and below it readily hydrolyzes. In soil or when diluted with water, metam-sodium is converted to methyl isothiocyanate (MITC). Other degradates of metam-sodium include carbon disulfide (CS2) and hydrogen sulfide (H2S). Metam-sodium is an agricultural general use pesticide used primarily as a broad spectrum preplant soil fumigant to control weeds, weed seeds, fungi, nematodes, and soil insects. End use products are formulated as 18–42% aqueous solutions sold under the trade names of Metam CLR, Vapam, and Sectagon. Metam-sodium has been registered since 1954. Registered uses of metam-sodium include agricultural soil fumigation, wood preservative, slimicide, tree-root killer, and aquatic weed control. Approximately 10 million pounds of metam-sodium were used in 1990, with 40–45% used for agricultural purposes (U.S. EPA, 1994a). As a soil fumigant, metam-sodium is applied after harvest and/or 14 to 21 days prior to planting by shank injection, disc, rotary tiller, drip irrigation, solid set sprinkler, or center pivot chemigation. In some parts of North America, fall applications are preferred because metam-sodium volatilizes over the winter and clears the soil, allowing planting to begin as soon as favorable springtime conditions arrive. By treating the soil with metam-sodium, fruit and vegetable Hayes’ Handbook of Pesticide Toxicology Copyright © 2001 Elsevier Inc. All rights reserved

growers can control weeds, reduce nematode populations, and control soil-borne pests. Metam-sodium may be used on all crops but is particularly important in the production of melons, peppers, tomatoes, potatoes, strawberries, citrus, grapes, almonds, artichokes, asparagus, carrots, lettuce, spinach, squash, forest tree seedlings, ornamentals, and cut flowers. By reducing competition from soil pests, metamsodium promotes healthier plants and increased yields. The U.S. EPA (1997) considers metam-sodium to be a commercially viable alternative to methyl bromide fumigation for fruit and vegetable production due to its low cost, wide range of control, and long record of safe use. It can be used to control weeds (e.g., bluegrass, Bermuda grass, chickweed, dandelion, ragweed, henbit, nutsedge, and wild morning glory), nematodes, and soil diseases caused by species of Rhizoctonia, Fusarium, Pythium, Phytophthora, Verticillium, and Sclerotinia (U.S. EPA, 1997). Metam-sodium has also been shown to be useful in integrated pest management systems as it can be used in conjunction with other treatment methods such as biological controls and soil pasteurization. Metam-sodium is a slightly to moderately toxic compound that when used according to label directions has been shown to be a safe and versatile product for over 45 years. For agricultural use, metam-sodium must be applied in a manner where there is no contact with workers or other persons, either directly or through drift. Only handlers equipped with the proper personal protection equipment may be in the area during application. In California, application must also be in compliance with the Technical Information Bulletin “Guidelines for All Application Methods for Metam-sodium in California.” The potential routes of human chemical exposures are oral (ingestion), dermal (direct skin contact), and inhalation, however, the chance for nonoccupational exposure to metam-sodium is minimal. Approved agricultural uses of metam-sodium do not leave residues on crops, thus eliminating diet as a source of exposure. The primary means of exposure to metam-sodium is through dermal occupational 2293

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exposure. Most of the potential for exposure to metamsodium itself comes from transloading and handling the liquid when preparing for application. The use of required protective gloves, boots, and clothing minimizes or eliminates dermal exposure to metam-sodium. The U.S. EPA’s Occupational and Residential Exposure Branch assumes that dermal exposure is minimal for handlers and nonexistent for nearby residents and bystanders (U.S. EPA, 1994a). There is little potential for inhalation exposure to metam-sodium, which has been proven through extensive monitoring and in a number of worker exposure studies. Proper protective equipment such as in-cab filtering systems and NIOSH-approved respirators that are used by workers provide protection in the unlikely situation that the liquid compound becomes aerosolized. The toxicology of “technical grade” or formulated metam-sodium (approximately 42% a.i.) is well established. Metam-sodium is synthesized in aqueous solution and then diluted with water, as necessary, to achieve the desired concentration and meet the label guarantee for the formulated product. Thus, “technical grade” is synonymous with the formulated material. Little toxicity information is available regarding pure or analytical-grade metam-sodium. The following discussion of technical grade/formulated metam-sodium toxicity briefly covers a number of published studies and the results of toxicity studies submitted to governmental agencies in support of metam-sodium registration.

107.2  Acute toxicity Metam-sodium is slightly to moderately acutely toxic depending on the route of exposure. The following toxicity values pertain to technical grade metam-sodium (U.S. EPA, 1994a). All of the following values were obtained with standard acute toxicity studies designed to determine the dose or concentration that causes death to 50% of the test animals (LD50 or LC50): The acute LD50 for technical grade metam-sodium (43.7% a.i.) is reported as 870 mg/kg for male rats and 924 mg/kg for female rats. The combined (male and female) LD50 is 896 mg/kg (placing the compound into Toxicity Category III (U.S. EPA, 1994a) or similarly classified as slightly toxic (LD50  5–15 g/kg; Klaassen, 1986). l The acute dermal LD50 of technical grade metamsodium (43.7%) applied to male and female rabbits is 368 mg/kg (Toxicity Category III). l The acute inhalation LC50 of aerosolized technical grade metam-sodium (42%) in rats is 2.275 mg/l (Toxicity Category III). l Technical grade metam-sodium (42%) was found to be slightly irritating to the eyes of New Zealand White rabbits (Toxicity Category III). l

Technical grade metam-sodium (42%) is irritating to the shaved skin of male and female rabbits (Liggett and McRae, 1991) and is classified as a moderate to severe dermal irritant (Toxicity Category II). l Metam-sodium (42%) was also found to be a skin sensitizer to guinea pigs using the delayed contact hypersensitivity test (Parcell and Denton, 1991). l

Acute studies conducted with 32.7% metam-sodium showed similar but milder results than the above cited data for the 42% formulated compound. Jowa (1998) reported the following values for multiple studies conducted with metam-sodium: The acute oral LD50 for 32.7% metam-sodium varied from 1294 to 1415 mg/kg for male rats and 1350 to 1428 mg/kg for female rats. l The acute dermal LD50 for 32.7% metam-sodium varied from 1012 to 3500 mg/kg in rabbits. l The acute inhalation LC50 varied from 4.7 to 5.4 mg/1 for male rats exposed to 32.7% metamsodium for four hours. l In one eye irritation study with rabbits, 32.7% metamsodium was found to be a mild irritant, but in another study it was found to be nonirritating. l Dermal irritation studies with rabbits exposed to 32.7% metam-sodium showed that the compound was a severe irritant in one study and was corrosive in another study. l Testing guinea pigs with 32.5% metam-sodium in the Buehler test resulted in sensitization. l

Standardized acute toxicity studies provide limited information regarding subtle toxic effects and are not designed to establish a no observed effect level (NOEL). To further understand the sublethal effects of a compound, lower dose levels or concentrations are required.

107.3  Subchronic toxicity Effects of metam-sodium exposure over longer periods vary with the species tested and route of administration. A variety of toxicity studies have shown that there is a definite dose–response effect to metam-sodium (U.S. EPA, 1992, 1993). At very low doses levels there is no evidence of toxicity, but as the dose level increases the prevalence and severity of toxic effects increases. In a 90-day study (Whiles, 1991), male and female mice were administered metam-sodium in drinking water at dose levels of 0, 0.018, 0.088, 0.35, or 0.62 mg/ml (2.7, 11.7, 52.4, or 78.7 mg/kg/day for males; 3.6, 15.2, 55.4, or 83.8 mg/kg/ day for females). No treatment-related mortality, morbundity, or clinical signs of toxicity were observed during the 90-day study period. Treatment-related statistically significant decreases in mean body weight were observed in both males and females at dose levels of 0.35 and 0.62 mg/ml. Treatment-related changes in hematology parameters

Chapter | 107  Metam-Sodium

were noted at doses as low as 0.088 mg/ml for females and 0.62 mg/ml for males. The lowest effect level was determined to be 0.088 mg/ml (11.7 mg/kg/day for males, 15.2 mg/kg/day for females) based on urinary bladder lesions observed in both males and females and in statistically significant decreases in hemoglobin, red blood cell, and hematocrit in females. The NOEL for systemic toxicity was 0.018 mg/ml (2.7 and 3.6 mg/kg/day for males and females, respectively). In another 90-day metam-sodium study (Allen, 1991), male and female rats received metam-sodium in the drinking water at nominal dose levels of 0, 0.018, 0.089, and 0.443 mg/ml (1.7, 8.1 and 26.9 mg/kg/day for males; 2.5,9.3, and 30.6 mg/kg/day for females). Systemic toxicity was evident by significant decreases in food and water consumption, decreased body weight gain, and histological changes in the nasal cavity olfactory epithelium in both males and females receiving metam-sodium at 0.443 mg/ml. Renal tubular dilation and basophilia along with increases in blood and protein in the urine were also observed in 0.443 mg/ml rats. In both males and females receiving 0.089 mg/ml there were significant decreases in red blood cell count and hematocrit. Females at the 0.089 mg/ml dose level also had a significant decrease in group mean body weight and decreased body weight gain (11%) when compared to controls. Based on the results of this study the NOEL was 0.018 mg/ml (1.7 mg/kg/day for males; 2.5 mg/kg/day for females). In a subchronic dog study (Brammer, 1992), metamsodium (43.15% purity) was administered by gelatin capsule to male and female beagles at nominal dose levels of 0, 1, 5, or 10 mg/kg/day once daily for 13 weeks. Toxic effects were observed at all dose levels tested but were primarily evident at the 5 and 10 mg/kg/day dose levels. Decreased body weight and body weight gain were observed in males and females receiving metam-sodium at 10 mg/kg/day. There were no significant clinical effects at 1 or 5 mg/kg/day and no ophthalmoscopic abnormalities in any animals. Regurgitation within 30–60 minutes of dosing occurred throughout the study in the 10 mg/kg/ day group and on isolated occasions in the 5 mg/kg/day dogs. There was no regurgitation in the 1 mg/kg/day dosing group. In dogs receiving 5 and 10 mg/kg/day there were changes in hematologic parameters (increases in cell volume, cell hemoglobin, neutrophils, and monocytes; decreases in mean corpuscular hemoglobin concentration); significant increases in plasma alanine aminotransferase (ALT), aspartate aminotransferase (AST), alkaline phosphatase (ALP), and gammaglutamyltransferase; increased blood, urobilinogen, bilirubin, and protein in the urine; and microscopic evidence of hepatitis). One female receiving 1 mg/kg/day showed increased plasma ALT. Biliary duct proliferation with inflammatory cell infiltration (less severe than hepatitis) was observed in one male and one female at the 5 mg/kg/day dose level and in one female at the 1 mg/kg/day dose level. No evidence of tumors were found in this study. Toxic effects appeared to be doseand time-related. For female dogs, no systemic NOEL was established (NOEL 1 mg/kg/day) due to increases in plasma

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ALT and biliary duct proliferation with inflammatory cell infiltration observed in a single female from the 1 mg/kg/day dose group. For male dogs, the systemic NOEL is 1 mg/kg/ day. The lowest observed effect level (LOEL) of 5 mg/kg/day is based on statistically significant increases in plasma ALT, AST, and alkaline phosphatase, and the increased incidence of hepatitis and bile duct proliferation. In order to further study the effects of metam-sodium on the liver of dogs, a study was conducted at the dose level that caused moderate to marked hepatitis in all dogs during the 90-day study described above (Brammer, 1993). One male and one female beagle dog received metam-sodium (43.14% purity) in a gelatin capsule daily at a dose level of 10 mg/kg. Dosing of each dog continued until there were elevations in plasma enzyme activities (or other clinical signs) indicative of liver toxicity. Following cessation of dosing, each dog was monitored until the enzyme activities returned to normal or prestudy levels. Dosing ceased after 12 weeks of dosing for the female and after 13 weeks for the male. Recovery was monitored for 8 weeks. After Week 6 the plasma ALT levels in the female began to increase and by Week 10 they were over 200 IU/L. In the male, elevated plasma ALT was noted at Week 9 and exceeded 200 UI/L by Week 11. In both dogs, plasma ALP levels gradually increased until dosing ceased. Following cessation of dosing, ALT levels increased during the first recovery week then gradually declined to normal levels after 8 weeks. Plasma ALP decreased in the female dog immediately after cessation of dosing and by Recovery Week 4 was less than prestudy values. In the male, ALP continued to rise during the first recovery week then gradually decreased so that by Recovery Week 5, ALP values were less than prestudy values. In both dogs, ALP levels continued to fall until study termination. At study termination, there were no macroscopic abnormalities in either dog and liver weights were normal. Microscopic evaluations revealed that there was a minimal or slight increase in the number of pigmented macropahges/Kupffer cells in the liver, but this is a common finding in beagle dogs of this strain (Alderley Park). The significant elevations in plasma ALT and ALP levels found in this study are consistent with the findings of the previous 90-day dog study at the same dose level (10 mg/kg/day) and are indicative of liver injury. However, after cessation of exposure, enzyme levels returned to normal, with full recovery eight weeks after the last exposure to metam-sodium. There was no evidence of liver injury at the end of the study. These findings confirm the reversible nature of induced liver effects from subchronic exposure to relatively high levels of metam-sodium.

107.4  Genetic toxicity Metam-sodium is not mutagenic but has been shown to be directly cytotoxic to bacteria, fungi, and mammalian cells. Metam-sodium has been tested and found to be negative in

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both in vitro and in vivo genetic toxicology assays covering a range of genetic toxicology endpoints including mutations, cytoge-netics, and DNA repair. There is evidence that at high enough dose levels, exposure to metam-sodium can be immunotoxic, with response evident in a dose-dependent manner. A review of metam-sodium genetic toxicity studies (Mackay, 1996) concluded that “[M]etam sodium shows no in vitro or in vivo genotoxic activity in a series of assays conducted up to concentrations/dose levels inducing significant toxicity in the target cells/animals.” In a bacterial gene mutation assay using Salmonella typhimurium strains TA92, TA98, TA100, TA1535, TA1537, and TA1538 in the presence and absence of metabolic activation (AROCHLOR 1254-induced rat liver S9 mix) there were no significant increases in the number of revertant colonies in any of the strains or S9 combinations tested. l In a Chinese hamster ovary mammalian cell gene mutation (HGRPT locus) assay metam-sodium was tested in the presence and absence of metabolic activation. There was no evidence of any reproducible dose-related effects of metam-sodium on mutation frequency or evidence of in vitro mutagenic activity. l In two in vitro cytogenetic assays using human lymphocytes there was no evidence of clastogenic activity from metam-sodium treatment when tested at concentrations up to those limited by toxicity and/or cytotoxic effects on chromosomal morphology. l The first study found an increase in aberrant cells at concentration levels that caused severe cytotoxicity (20 g/ml without S9 mix; 40 and 20 g/ml with S9 mix) and therefore were unsuitable to be included in the evaluation of clastogenic potential. At concentration levels of 1, 5, and 10 (g/ml with and without S9 mix, there were no increases in the percentage of aberrant cells. l In the second in vitro human lymphocyte clastogenic study, metam-sodium at concentrations of 2.5, 20, and 30 g/ml in the absence of S9 mix and 5, 20, and 40 (g/ml with S9 were tested. No statistically or biologically significant increases in the percentage of aberrant cells were observed at any of the metamsodium concentrations tested in the absence of the S 9 mix. There was a small statistical increase in the number of aberrations observed in the 40 g/ml test concentration with the S9 mix, but the values observed were well within the historical solvent control range and do not indicate clastogenic activity. l In an in vitro unscheduled DNA synthesis assay using primary rat hepatocytes treated with metam-sodium there was no evidence of induction of DNA repair, even in cultures treated with toxic concentrations of metam-sodium. l In an in vivo Chinese hamster bone marrow chromosomal aberration assay there was no evidence of any l

polyploidy inducing effect of metam-sodium nor was there evidence of any clastogenic activity. l When metam-sodium was administered to CD-1 mice in an in vivo mouse bone marrow micronucleus test, there was no evidence of clastogenic activity in the mouse bone marrow when tested up to the maximum tolerated dose level for both male and female mice. There were no statistically or biologically significant increases in the incidence of micronucleated polychromatic erythrocytes. U.S. EPA Tox Oneliners report on the results of genetic studies submitted to and reviewed by the U.S. EPA. Jowa (1998) reported on the same genetic studies submitted to and reviewed by the California Environmental Protection Agency (Cal EPA). In some cases, the results presented by Jowa did not agree with conclusions of the U.S. EPA. In two separate Ames studies using multiple strains of Salmonella typhimurium (TA 1535, 1537, 1538, 92, 98, and 100) up to 2500 g/plate with and without activation (S9 mix), metam-sodium did not induce mutations and the results were negative. l In a study with yeast (Sacchromyces cerevisiae strain D4) with and without S9 mix, metam-sodium did not induce mutations and the results were negative. l In an in vitro study with Bacillus subtilis, metamsodium did not cause DNA damage. l According to the Cal EPA review, equivocal results were obtained for a REC assay in Bacillus subtilis H17 and M45 (/ S9). l According to the U.S. EPA review, metam-sodium (42.2%) is not a recombinogenic agent (i.e., causes DNA damage) to Bacillus subtilis strains H17 and M45 at concentrations up to 150 l/well. l In an in vitro study using cultured lymphocytes procured from a single male human donor, there was evidence of possible aberrant chromosomes. However, according to the Metam-sodium Task Force, the scientific validity of this study is under question since the aberrant chromosomes were observed only at concentration levels that were clearly cytotoxic to the cells. When the cells from noncytotoxic concentration levels were evaluated, there was no indication of any clastogenic activity. l In a mammalian cytogenetic study with Chinese hamsters, metam-sodium did not induce cytogenic effects l According to the Cal EPA review, there was evidence of polyploidy in Chinese hamster ovary cells at dose levels of 150 and 300 mg/kg. l According to the US EPA review, metam-sodium (42.2%) had a negative response (no effect) in the Chinese hamster bone marrow cytogenetic assay at concentrations of 150, 300, and 600 mg/kg. l Metam-sodium was found to be negative in an unscheduled DNA synthesis study with primary rat hepatocyte culture. l

Chapter | 107  Metam-Sodium

A study was conducted to assess the immunotoxicological and selected general toxicological effects of metam-sodium (Pruett et al., 1992). Metam-sodium was administered to female B6C3F1 mice at 200 mg/kg/day for 3, 5, 10, or 14 days. Selected organ weights were measured, hematological and bone parameters were examined, changes in thymus and spleen lymphocyte subpopulations were evaluated, and production of antibody-forming cells in vitro was measured. Major effects of metam-sodium administration included decreased thymus weight at all time points; increased spleen weight and bone marrow cellularity after 10 or 14 days of exposure; significant decreases in mature lymphocytes in the thymus and spleen; decrease in thymocytes; and decreased body weight. According to Pruett and co-workers (1992), overall patterns of change indicate that metam-sodium rapidly depletes most CD4  CD8  thymocytes, more slowly depletes a smaller number of mature lymphocytes in the thymus and spleen, and induces compensatory and/or detoxication mechanisms after 10–14 days of exposure. Pruett and co-workers (1992) conducted subsequent experiments to assess selected immune function parameters after exposure to metam-sodium. Metam-sodium was administered for seven days (either orally or dermally) and immunological assays were conducted on Day 8. Mice receiving metam-sodium orally at dose levels of 50 to 300 mg/kg showed substantial, dose-dependent suppression of NK cell activity. Evaluation of humoral responses indicated that the cellular and molecular components required for humoral immune responses are not major targets for the acute effects of metam-sodium. There was no suppression of antibody production in vivo or splenocyte responses to mitogens or allogeneic lymphocytes in vitro, which indicates that the lymphocytes which survive metam-sodium exposure are still able to proliferate and differentiate and are not significantly impaired with regard to function. The authors also noted that the pattern of thymic subpopulation changes is consistent with direct or indirect induction of apoptosis. These studies showed that immunological parameters could be significantly suppressed in the absence of a significant decrease in body weight, suggesting that most of the effects of metam-sodium on the immune system are not secondary to generalized toxicity. In response to reports that metam-sodium is immunotoxic, a series of in vivo and in vitro studies was conducted with metam-sodium and other dithiocarbamates (Padgett et al., 1992). Metam-sodium in distilled water was administered orally via daily gavage to female mice for seven days at dose levels of 0, 150, 225, or 300 mg/kg. Body weight was not significantly decreased at any dose level, but thymus weight was significantly decreased in mice receiving metam-sodium at dose levels of 225 and 300 mg/ kg. In tests of splenic NK cell activity, metam-sodium at dose levels of 225 and 300 mg/kg was found to significantly inhibit NK activity. This study also demonstrated

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that metam-sodium was directly cytotoxic to lymphoid cells in vitro, but that cytotoxic potency in vitro does not correlate well with immunological changes in vivo.

107.5  Developmental and reproductive toxicity Developmental studies in two different species found evidence of increased fetal loss, increased skeletal variations, and developmental delays from oral administration of metam-sodium to pregnant animals at dose levels that also caused overt maternal toxicity. Visceral or skeletal abnormalities were not present at low dose levels but increased in incidence and severity with increasing dose (U.S. EPA, 1991). In a multigeneration reproductive study, metamsodium did not affect reproductive performance, even at toxic dose levels (U.S. EPA, 1994a). A developmental study with rats receiving metamsodium (Hellwig and Hildebrand, 1987) indicated that there were significant maternal and fetal effects at higher dose levels and that these effects were dose-related. An aqueous solution of metam-sodium (42.2%) was administered at 0, 10, 40, or 120 mg/kg by gavage to pregnant Wistar rats on Days 6–15 of gestation. Body weight gains were significantly decreased in dams receiving metam-sodium at dose levels of 40 and 120 mg/kg during the dosing period. Cesarean section observations revealed that there was a statistically significant increase in the percentage of postimplantation loss and a significant decrease in the percentage of live fetuses per dam at the 10 and 120 mg/kg dose levels, but not at the 40 mg/kg dose level. It is possible that the effects observed at the 10 mg/kg dose level were statistical anomalies, but this remains unconfirmed in the absence of a review of the individual data, which were not available. All other parameters in the 10 mg/kg group were comparable to controls, including the total number of live fetuses and live fetuses per dam. Since there were no statistically significant changes in Cesarean section observations in the 40 mg/kg group, it is likely that the statistically significant changes in percentage of live fetuses per dam and the percentage of postimplantation loss in the 10 mg/kg group are not treatment-related. The only abnormal finding observed during the macroscopic examination of the fetuses was meningocele (hernial protrusion of the meniges through a bony defect) in two fetuses from one litter in the 120 mg/kg dose group. Since this is a rare finding that was not present in historical controls, this anomaly was considered to be treatment-related. Skeletal evaluations of the fetuses revealed an increased incidence of variations and a delay in the development of fetuses in the 40 and 120 mg/kg dose groups. Fetal weights were significantly reduced in the 120 mg/kg group. The NOEL for fetal and maternal effects was 10 mg/kg. In another rat developmental toxicity study (Tinston, 1993) groups of pregnant rats were administered

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metam-sodium at dose levels of 0, 5, 20, or 60 mg/kg/day on Days 7–16 (inclusive) of gestation. Maternal toxicity evidenced by reduced body weight gain, reduced food consumption, and the presence of clinical signs (piloerection, salivation, and urinary incontinence) occurred at the 20 and 60 mg/kg/day dose levels. Body weight gain and food consumption were marginally reduced at the 5 mg/kg/day dose level but there were no treatment-related clinical signs. In both the 20 and 60 mg/kg/day dose groups there was an increase in fetal effects (reduced fetal weights, reduced ossification of manus and pes, and increased incidences of minor skeletal defects and/or variants). The no observed adverse effect level (NOAEL) for maternal toxicity or fetal effects in this study was 5 mg/kg/day. In a teratology/developmental study (Hellwig, 1987), pregnant Himalayan rabbits were administered a 42.2% aqueous solution of metam-sodium at dose levels of 0, 10, 30, or 100 mg/kg by gavage from gestation Days 6 through 18. Evaluation of body weight data revealed a treatment-related decrease in body weight gain in the 100 mg/kg dams. There were no statistically significant treatment-related effects noted in food consumption or food efficiency. Cesarean section observations revealed statistically significant decreases in the total number of live fetuses and statistically significant increases in total re-sorptions in the 30 and 100 mg/kg/day groups. Macroscopic fetal examinations revealed meningocele and spina bifida in one rabbit in one litter in the 100 mg/kg/day group (it is not clear from the data if the findings were in the same rabbit or in two separate rabbits). Due to the rarity of this event and that it was also present in the rat developmental study, this abnormality is considered to be treatment-related. There were no treatment-related effects noted from the visceral examinations. Skeletal examinations revealed no treatment-related effects. However, these examinations were done using acceptable European methods that have not been validated by EPA and are not considered to be comparable with U.S. EPA-accepted methods. In another rabbit developmental toxicity study (Hodge, 1993) groups of pregnant rabbits were administered metam-sodium at dose levels of 0, 5, 20, or 60 mg/kg/day on Days 8–20 (inclusive) of gestation. At the 60 mg/kg/day dose level, dams showed marked weight loss and reduced food consumption. At 20 mg/kg/day, body weight of dams was slightly reduced. There were no observable effects noted on dams at the 5 mg/kg/day dose level. Fetal examinations revealed a marked increase in embryonic lethality at the 60 mg/kg/day maternal dose level and changes in ossification pattern at the 20 and 60 mg/kg/day maternal dose levels. The NOAEL for maternal and developmental toxicity was 5 mg/kg/day. In a multigeneration reproduction study (Milburn, 1993), Alpk:ApfSD rats received metam-sodium in drinking water at the following concentrations: 0,0.01,0.03, or 0.1 mg/ml. These concentrations corresponded to dose levels of 0, 1.2, 3.2, or 11.5 mg/kg/day for males and 0, 1.8, 3.9,

Hayes’ Handbook of Pesticide Toxicology

or 13.5 mg/kg/day for females. After the first 10 weeks of treatment, animals were mated on a one-to-one ratio. Males were then removed from their cages and females were allowed to give birth and raise pups. At 21 days of age, pups from the parental (F0) generation were selected as parents for the Fl generation. In parents, body weights were marginally reduced in rats receiving 0.10 mg/ml (the highest concentration tested) during the premating period and markedly reduced during pregnancy and lactation. Water consumption was reduced in the 0.10 mg/ml rats throughout the study and to a lesser extent in the 0.03 mg/ml group. In offspring, there was a marginal reduction in food consumption during the premating period in the F0 and Fl rats in the 0.10 mg/ml group, but there were no effects on food consumption in the 0.01 and 0.03 mg/ml treatment groups. Offspring body weights and total litter weights were reduced in the 0.10 mg/ml group in both generations. There were no effects on any of the reproductive parameters at any treatment level for parents or offspring. Histopathological evaluations indicated increased changes in the epithelium of the nasal passages of the F0 and Fl adult females in the 0.10 mg/ml groups. This effect was not observed in 0.10 mg/ml adult males or in male or female offspring of either generation. No treatment-related histopathological changes were observed in rats receiving metam-sodium at concentrations of 0.01 or 0.03 mg/ml. Metam-sodium did not affect reproductive performance at any dose level tested. Evidence of toxicity was observed only at the highest concentration level tested, i.e., 0.1 mg/ml. In adult female rats receiving metam-sodium at the 0.1 mg/ml concentration level (13.5 mg/kg/day), evidence of systemic toxicity consisted of (1) duct hypertrophy of Bowman’s gland with loss of alveolar cells, (2) degeneration, disorganization, and/or atrophy of the olfactory epithelium, and (3) dilation of the Bowman’s gland ducts. Changes in Bowman’s glands were accompanied in all affected animals by degeneration, disorganization, and/or atrophy of the olfactory epithelium. In pups in the 0.1 mg/ml group, evidence of toxicity consisted of a 14% decrease in mean pup weight on Day 22 for the F1 generation, a 16% decrease in mean body weight gain for F2 litters, and decreases of 8–9% in testes and epididymis weight in male pups in the F1a and F2a litters. The NOEL for systemic toxicity (adults and pups) was 0.03 mg/ml. The NOEL for reproductive effects was 0.1 mg/ml (11.5 mg/kg/day for males, 13.5 mg/kg/day for females).

107.6  Chronic/oncogenicity toxicity A two-year combined chronic toxicity/carcinogenicity study demonstrated that metam-sodium shows no carcinogenic potential in rats (Thomassen, 1998; U.S. EPA, 1994b). However, a two-year carcinogenicity study in mice

Chapter | 107  Metam-Sodium

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Table 107.1  Summary of Metam-Sodium No Observed Effect Levels (NOELs) and Lowest Observed Effect Levels (LOELs) Study

Species

Dosing duration

Dose levels

NOEL

LOEL effects

90-day drinking water

Mouse

90 days

0, 0.018, 0.088, 0.35, and 0.62 mg/ml

0.018 mg/ml

0.088 mg/ml

0, 2.7, 11.7, 52.4, and 78.7 mg/kg/day ( ) 0, 3.6, 15.2, 55.4, and 83.8 mg/kg/day ( )

2.7 mg/kg/day

0,0.018, 0.089, and 0.443 mg/ml

0.018 mg/ml

  urinary bladder lesions  decreases in hemoglobin, RBC, and hematocrit

●

90-day drinking water

Rat

90 days

●

3.6 mg/kg/day 0.089 mg/ml  decreased body weight and body weight gain

●

90-day oral

Dog

90 days

0, 1.7, 8.1, and 26.9 mg/kg/day ( ) 0, 2.5, 9.3, and 30.6 mg/kg/day ( )

1.7 mg/kg/day

0, 1, 5, and 10 mg/kg/day

1 mg/kg/day

2.5 mg/kg/day

  decreases in RBC and hematocrit

●

5 mg/kg/day  increased plasma ALT, AST, and ALP ●  hepatitis and bile duct proliferation ●

1 mg/kg/ day

1 mg/kg/day   increased plasma ALT   bile duct proliferation

● ●

Developmental toxicity

Rat

10 days

0, 10, 40, and 120 mg/kg/day

10 mg/kg/day

40 mg/kg/day ●  decreased maternal weight gain ●  increased fetal skeletal variations ●  delay in development

Developmental toxicity

Rat

10 days

0, 5, 20, and 60 mg/kg/day

5 mg/kg/day

20 mg/kg/day ●  decreased maternal weight gain ●  reduced food consumption ●  maternal clinical signs ●  increased fetal skeletal variations ●  reduced ossification of manus and pes ●  reduced fetal weights

Developmental toxicity

Rabbit

13 days

0, 10, 30, and 100 mg/kg/day

10 mg/kg/ day—fetal

30 mg/kg/day—fetal ●  decreases in live fetuses ●  increased resorptions 100 mg/kg/day—maternal ●  decreased body weight gain

30 mg/kg/ day—maternal Developmental toxicity

Rabbit

13 days

0, 5, 20, and 60 mg/kg/day

5 mg/kg/day

20 mg/kg/day ●  reduced maternal body weights ●  change in fetal ossification pattern

Multigeneration

Rat

Chronic

0,0.01, 0.03, and 0.1 mg/ml

0.03 mg/ml— systemic

0.1 mg/ml—systemic toxicity

0, 1.2, 3.2, and 11.5 mg/ kg/day ( )

 wchanges in Bowman’s gland and olfactory epithelium (adults)

●

(Continued)

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Table 107.1  (Continued) Study

Species

Dosing duration

Dose levels

NOEL

0, 1.8, 3.9, and 13.5 mg/ kg/day ( )

LOEL effects  decreased mean pup weight   0.1 mg/ml ●

0.1 mg/ml— reproductive Carcinogenicity: twoyear drinking

Rat

Chronic

0, 0.019, 0.056, and 0.19 mg/ml 0, 1.3, 3.9, and 12.0 mg/ kg/day ( ) 0, 2.3, 6.2, and 16.2 mg/ kg/day ( )

0.056 mg/ml

0.19 mg/ml ●  decreased body weight gain ●  decreased food consumption, food efficiency and water consumption ●  changes in hematology and clinical chemistry ●  abnormalities in nasal cavity, voluntary muscle and sciatic nerve

Carcinogenicity: twoyear drinking

Mouse

Chronic

0, 0.019, 0.074, and 0.23 mg/ml 0, 1.6, 6.5, and 27.7 mg/ kg/day ( ) 0, 2.3, 8.7, and 29.9 mg/ kg/day ( )

0.019 mg/ml

0.074 mg/ml ●  increased liver weight ●  changes in kidney and epididymis weights

1-year oral

Dog

Chronic

0, 0.05, 0.1, and 1.0 mg/kg/day

0.1 mg/kg/day

1.0 mg/kg/day ●  increase in hepatocyte and liver macrophage/Kupffer cells ●  increased plasma ALT

Acute neurotoxicity

Rat

Single dose

0, 22, 324, and 647 mg/kg

22 mg/kg

22 mg/kg ●  reduced ambulatory and total motor activity

Subchronic neurotoxicty

Rat

13 weeks

0, 0.02, 0.06, and 0.2 mg/ml

0.06 mg/ml ( )

0, 1.4, 5.0, and 12.8 mg/ kg/day ( ) 0, 2.3, 7.0, and 15.5 mg/ kg/day ( )

0.02 mg/ml ( )

0.2 mg/ml ( ) 0.06 mg/ml ( ) ●  decreased body weight gain

revealed an increased incidence of angiosarcoma in mice at higher dose levels (U.S. EPA, 1994c). A one-year study with dogs showed no evidence of carcinogenicity but evidence of liver damage similar to but less severe than the effects (that were shown to be reversible) observed in previous subchronic metam-sodium dog studies. In a two-year combined chronic toxicity/carcinogenicity study with Wistar rats (Rattray, 1994), metamsodium (43.14% a.i.) was administered in drinking water at concentration levels of 0, 0.019, 0.056, or 0.19 mg/ml (achieved dosages of 0, 1.3, 3.9, or 12.0 mg/kg/day for males and 0, 2.3, 6.2, or 16.2 mg/kg/day for females). There was no evidence of an adverse effect of metamsodium on the survival or rats. There were no ophthalmological changes associated with metam-sodium treatment. Evidence of toxicity was present in both males and females at the highest concentration level tested, i.e.,

0.19 mg/ml. At 0.19 mg/ml, male and female rats had decreased mean body weight gain for Weeks 1–13 (12% for males, 16% for females) and for Weeks 1–105 (18% for males, 20% for females). Food consumption, food efficiency, and water consumption were significantly decreased for both males and females receiving 0.19 mg/ml metamsodium. Effects were also observed in 0.19 mg/ml male and female hematology (decreased red blood cells, hemoglobin, and hematocrit) and clinical chemistry (decreased cholesterol and triglycerides). Nasal passages were identified as the target organ. Microscopic abnormalities of the nasal cavity were mainly confined to 0.19 mg/ml animals. These changes included (1) an increased incidence of rhinitis, (2) hypertrophy of Bowman’s ducts/glands, (3) atrophy and adenitis of Steno’s gland, and (4) hyperplasia and degeneration of olfactory epithelium. The incidence of degenerative myopathy of voluntary muscle was similar in all

Chapter | 107  Metam-Sodium

groups, including controls. However, there was an increase in the severity of myopathy in animals in the 0.019 mg/ml group. There was no indication of an increased incidence of neoplasia or early onset of tumors from treatment with metam-sodium. Evaluation of the tumor incidence demonstrated that metam-sodium shows no carcinogenic potential in rats (Rattray, 1994; U.S. EPA, 1994b). The NOEL for both male and female Wistar rats was 0.056 mg/ml. The Cal EPA, Department of Pesticide Regulation evaluated the tumor data from the two-year metam-sodium drinking-water study in rats and concluded that there was a possible tumorigenic effect at the 0.056 mg/ml concentration level (U.S. EPA, 1995). According to the Cal EPA review, the incidence of hemangiosarcoma (8/64) was increased at this dose, in relation to the control incidence (0/64) and the high dose (0.19 mg/ml) incidence (3/64). The hypothesis that this could be a positive response was based on the positive findings in the two-year mouse study and that this increased incidence could be based on decreased body weight observed at the high dose in relation to other doses. When the U.S. EPA (1995) re-evaluated the tumor data for their Carcinogenicity Classification, they did not find the effect that the Cal EPA found (presumably because the Cal EPA analysis did not exclude animals that died before observation of the first tumor). However, there was a significant pairwise comparison in the incidence of hemangiosarcoma in male rats at the 0.019 and 0.056 mg/ml (1.3 and 3.9 mg/kg/day) levels when compared to controls. The U.S. EPA also considered debatable the hypothesis of increased incidence of hemangiosarcoma at the mid-dose level based on decreased body weight in male rats at the high dose level. Rats in this study were not fed a calorie-restricted diet, nor was their access to food controlled. In addition, the decreases in body weight gain were observed for both male and female rats, although the preponderance of hemangiomas/hemangiosarcomas was observed only in male rats. In addition, the time to tumor formation was observed at approximately the same time in all dose levels. In calorierestricted studies, the numbers of tumors are often reduced in conjunction with a delay in the time to tumor formation. In response to the position taken by Cal EPA, the two-year drinking water study with Wistar rats was reviewed and compared to an expanded historical control data base for Wistar rats that at the time of their original review was not available (Thomassen, 1998). Hemangiomatous tumors (heman-gioma and/or hemangiosarcoma) were observed only in rats sacrificed at termination of the study (Study Week 105) or in rats that were found dead or were euthanized due to their clinical condition (moribund or to prevent suffering). No hemangiomatous tumors were observed in rats euthanized during the interim sacrifice at Study Week 53. Therefore, tumor analysis (as was done by the U.S. EPA) should exclude animals that were sacrificed or died prior to the first observance of a heman-gioma or hemangiosarcoma. Statistically significant increased incidence of hemangiosarcomas occurred only in

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the 0.019 and 0.056 mg/ml males and not in the 0.19 mg/ml males, although the actual numbers were very similar (3/49 at 0.019 mg/ml and 3/51 at 0.19 mg/ml). Three possible explanations for reduced tumor incidence with an increase in treatment are: (1) the high dose of metam-sodium exceeded the maximum tolerated dose and had a negative impact on the tumor response in the high dose males; (2) reduced body weight associated with reduced tumor incidence accounted for the difference (as suggested by the Cal EPA reviewer); or (3) biological variability was responsible. Neither the U.S. EPA nor the Cal EPA thought the maximum tolerated dose had been exceeded. The U.S. EPA carcinogenicity peer review panel did not believe that reduced body weight accounted for the reduced tumor incidence (U.S. EPA, 1995). However, the possibility that biological variability could account for the effect was hampered by the lack of historical control data for Wistar rats. Hemangiomatous tumors (variously diagnosed as angiomas and angiosarcomas, hemangiomas and hemangiosarcomas, and lymphangiomas and lymphangiosarcomas) are common in some but not all strains of Wistar rats (Bomhard, 1992; Bomhard et al., 1986; Crain, 1958; Deerberg et al., 1980; Kroes et al., 1981; Rehm et al., 1984). The reported incidences of these tumors in Wistarderived rats used in European laboratories vary considerably, but there are reports of up to a 74% incidence for males and 44% for females (Rehm et al., 1984). These reports also indicate that there is a definite propensity for development of tumors in the lymph nodes, particularly the mesenteric lymph nodes of male Wistar rats. Although Zeneca Central Toxicology Laboratory did not have an historical control data base for Wistar rats used in this study, there was a large historical control tumor data base compiled by several European laboratories utilizing Wistarderived rats (49 studies ranging in duration from 24 to 31 months). This data base was published as the RITA Wistar Rat Control Tumor Data Base (Thomassen, 1998). Information presented in the RITA control data base is consistent with the types and numbers of tumors observed in the metam-sodium two-year rat study. Based on a thorough review of the original study and comparisons with the RITA control data base for Wistar rats, it was concluded that: Metam-sodium is not a carcinogen in the rat. The reduced number of hemangiosarcomas in the high dose male rats in the metam-sodium study is not due to reduced caloric intake. l The natural distribution and incidence of spontaneously occurring hemangiosarcomas in untreated male Wistar rats can account for the distribution and incidence of hemangiosarcomas observed in male rats treated with metam-sodium. l l

In a two-year carcinogenicity study in mice (Horner, 1994), metam-sodium (43.15%) was administered in the drinking water to C57BL/10JfCD-l/Alpk mice for 104 weeks at

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nominal concentration levels of 0, 0.019, 0.074, or 0.23 mg/ml (actual achieved doses of 0, 1.6, 6.5, or 27.7 mg/kg/day for males and 0, 2.3, 8.7, or 29.9 mg/kg/day for females). Metamsodium did not adversely affect survival of mice at any dose level. Clinical signs of toxicity were considered to be unremarkable. Male and female mice receiving metam-sodium at 0.074 and 0.23 mg/ml had dose-related and statistically significant increases in absolute liver weight when compared to controls (111% and 119% for 0.074 mg/ml males and females, respectively; 135% and 122% for 0.23 mg/ml mice). At 0.23 mg/ml, male mice had decreased body weight gain of 14% for Weeks 1–13 and 20% for Weeks 1–104. Food consumption was unaffected during the early part of the study, but during Weeks 24 to 52 there were statistically significant decreases in food consumption for the 0.074 and 0.23 mg/ml male mice. Decreases in food consumption were not observed for female mice. Water consumption was significantly decreased for both males and females in the 0.23 mg/ml group during the study’s first week, but by Week 9, males in the 0.23 mg/ml group had significantly increased water consumption. By Week 11, water consumption was significantly increased for both 0.074 and 0.23 mg/ml males. By Week 48, water consumption for all groups (male and female) were approximately equal to controls. Hematological investigations showed no significant treatment-related effects at any dose level. Macroscopic observations revealed several changes in 0.23 mg/ml mice including liver appearance (accentuated lobular pattern, pale), subcutaneous tissue masses, urinary bladder wall thickening, and reduced incidence of enlarged seminal vesicles. Several changes were noted in liver, kidney, and epididymis weights at 0.074 and 0.23 mg/ml treatment levels. Microscopic evaluations revealed several non-neoplastic effects in 0.23 mg/ml mice but also revealed evidence of neoplastic changes at this same dose level. There was evidence of dose-dependent metam-sodium induced carcinogenicity in mice. In both males and females at the 0.23 mg/ml treatment level, there was an increased incidence of hepatic adenoma and angiosarcoma, splenic angiosarcoma, subcutaneous tissue angiosarcoma, and a single incidence of a urinary bladder transitional cell papilloma in one high dose male and a single incidence of urinary bladder transitional cell carcinoma in one high dose female. The overall incidence of angiosarcoma, regardless of site, increased for both males and females in the 0.23 mg/ml treatment group when compared to concurrent as well as historical controls. The no observed effect level for neoplastic changes is 0.074 mg/ml. According to the U.S. EPA (1994c), there was equivocal evidence of a possible increase in splenic angiosarcoma at 0.074 mg/ml [something that the study author and registrants believe is related to the difficulty in determining the primary site(s) of angiosarcoma]. There was no evidence of increased tumors at the lowest dose level. In the U.S. EPA’s (1994c) evaluation of the two-year mouse study, the reviewers suggested that the dosing levels in this study

Hayes’ Handbook of Pesticide Toxicology

were adequate due to the degree of toxicity (increased liver weights, non-neoplastic changes in bladder, and tumors) observed in both males and females at the 0.23 mg/ml treatment level. According to the U.S. EPA (1994c) based on the significant increase observed in liver weight in male and female mice, the LOEL is considered to be 0.074 mg/ml, which is the NOAEL for neoplastic changes. In a one-year toxicity study (Brammer, 1994), metamsodium was administered orally to beagle dogs at dose levels of 0, 0.05, 0.1, or 1.0 mg/kg/day. Animals were observed daily for food consumption, evidence of gastro-intestinal upset, and changes in clinical condition. Animals also received detailed clinical evaluations weekly and complete veterinary examinations (including ophthalmoscopy) every three months. Blood chemistry, hematology, urine chemistry, and cytology evaluations were conducted at regular intervals throughout the study. At study termination, each animal received a full necropsy and histopathological evaluation of selected tissues. Throughout the study, there were no overt signs of toxicity at any dose level and all dogs remained in good health. There were no toxicologically significant effects on body weight, food consumption, clinical condition, or on the incidence of gastro-intestinal effects (i.e., vomiting, loose stools, etc.). There were no ophthal-moscopic abnormalities nor were there significant changes in hematology or urinalysis or in organ weights. There were no macroscopic findings that could be attributed to treatment with metamsodium. Microscopic evaluations revealed a slight increase in hepatocyte and macrophage/Kupffer cells in the liver of one female dog dosed at 1.0 mg/kg/day. This same female also had significant elevations in plasma alanine transaminase activity. These changes were similar to but less severe than those observed in previous subchronic dog studies with metam-sodium and are considered to be treatment-related. Therefore, the NOEL for this study was 0.1 mg/kg/day.

107.7  Nurotoxicity Metam-sodium is not neurotoxic based on evidence from neurotoxicity studies. In an acute neurotoxicity study (Lamb, 1993), male and female Sprague–Dawley Crl: CD®BR rats received metamsodium (43.15%) orally at doses of 0, 50, 750, or 1500 mg formulated metam-sodium/kg body weight or 0, 22, 324, or 647 mg a.i./kg. Mortality was observed at the l500 mg/kg dose level (males 31%, females 19%). Signs of systemic toxicity were observed at the 750 and 1500 mg/kg dose levels and included changes in posture, palpebral closure, respiratory rate, arousal, rearing activity, time to first step, olfactory and pupil responses, tail pinch response, hindlimb strength, body temperature, and body weight. Lacrimation and salivation were also noted among some animals at both the 750 and 1500 mg/kg dose levels. Reductions in ambulatory and motor activity were observed at the 50 mg/kg dose level and above

Chapter | 107  Metam-Sodium

on Day 0 (day of dosing) yet there were no treatment-related effects on the functional observational battery in the 50 mg/kg dose group. No signs indicative of neurotoxicity were observed at any dose level. There was no significant change in brain cholinesterase (ChE) activity at any dose level and there were no signs of cholinergic effects at any dose level. There were no treatment related differences in brain weight or dimensions in any treatment group. Histopathological evaluations of brain and nervous system tissues showed no evidence of neurotoxicity. According to the U.S. EPA (1994d), plasma and RBC ChE activity levels were reduced in 1500 mg/kg male and female rats 24 hours postdose (6% and 12% for male plasma and RBC ChE, respectively; 24% and 14% for females). [Although statistically significant, none of these decreases in ChE activity are considered to be biologically relevant as all decreases are well within the range of normal variation and are below the thresholds set by the World Health Organization (JMPR, 1995; WHO, 1990) and other regulatory agencies (Carlock et al., 1999).] Based on the results of this study, the 1500 mg/kg dose level was considered the NOAEL in males and females for acute neurotoxicity while the 50 mg/kg/day dose level was considered the LOEL for acute systemic toxicity (based on reduced motor activity). In a subchronic neurotoxicity study (Allen, 1991), male and female Sprague-Dawley rats were given metam-sodium (43.15%) in drinking water at concentration levels of 0, 0.02, 0.06, or 0.2 mg/ml for 13 weeks (achieved dosages of 0, 1.4, 5.0, or 12.8 mg/kg/day for males and 0, 2.3, 7.0, and 15.5 mg/ kg/day for females). Male and female rats administered 0.2 mg metam-sodium/ml drinking water showed reductions in body weight, food consumption, and water consumption. Similar effects were observed in females at the 0.06 mg/ml concentration level. Body weight gain was reduced 14% for the 0.2 mg/ ml males and the 0.06 mg/ml females, and 18–21% for the 0.2 mg/ml females. Food utilization was slightly reduced in males at the 0.2 mg/ml level. Reduced water consumption was also observed in males at the 0.06 mg/ml level and in females at the 0.02 mg/ml level. All of these effects were considered to be a consequence of poor potability of the drinking water rather than toxicity of metam-sodium. A functional observational battery and comprehensive neuropathological examination of the peripheral and central nervous systems revealed no evidence of any effects attributable to treatment with metam-sodium. Since there was no evidence of a neurotoxic effect from metamsodium, the NOAEL for neurotoxicity is 0.2 mg/ml.

107.8  Other studies (Mammalian) In vitro percutaneous absorption of metam-sodium through rat and human skin was evaluated (Clowes, 1993). Metamsodium was applied at dose levels of 940 and 94.0 g/cm2.

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Ten hours after dermal application, the skins were washed to determine how much of the dose could be removed from the skin surface, receptor fluid was analyzed, and the proportion of the dose remaining associated with the skin after washing and the amount absorbed were quantified. The absorption of metam-sodium was found to be dose and time dependent through both rat and human skin. The highest amount of metam-sodium absorbed was through rat skin from the 940 g/cm2 application (mean 200 g/cm2; 21.3% of the applied dose at 10 hours). A correspondingly smaller amount was absorbed from the 94.0 g/cm2 application through rat skin (mean 18.2 g/cm2; 19.4% at 10 hours). Absorption through cadaver human skin was 2.19% for the 940 g/cm2 dose (mean 20.6 g/cm2 ) and 12.2% (mean 11.5 g/cm2) of the applied dose. Absorption of metamsodium through both rat and human skin increased with time but at a decreasing rate over the 10 hour period. The percentage of the dose remaining in the skin increased with decreasing dose. There was less metam-sodium absorbed through human skin than rat skin, and at the highest dose level there was approximately a 10-fold decrease in absorption of metam-sodium by human skin when compared to rat skin. An in vivo percutaneous (dermal) absorption study in the rat (Stewart, 1992) showed that metam-sodium and/ or its radiolabeled degradation products are only poorly absorbed following a single dermal application to the rat. Radiolabeled 14C metam-sodium was applied to male rats in aqueous solution at nominal dose levels of 0.1, 1, and 10 mg/animals. A glass saddle containing an activated charcoal filter to adsorb any volatile radioactivity evaporating from the skin surface protected the application site. Four animals from each group were evaluated at 1, 2, 10, and 24 hours after treatment for radioactivity in the excrement, in and on the skin, and in the body. Another four animals per group had the treatment area washed 10 hours after administration. Radioactivity in the excrement was monitored over a total of 72 hours prior to evaluations of the skin and body. Overall mean recoveries of radioactivity were in the range of 83.5 to 95.7% of the applied dose. The extent of absorption was similar at each dose level with an overall mean of approximately 3%. In general, absorption increased with time. Levels of absorbed material 24 hours postapplication for the 0.1, 1, and 10 mg/animal dose levels were approximately 7.5, 50, and 231 g equivalents of 14C metam-sodium, respectively. Substantial quantities of the nonabsorbed dose were recovered from the charcoal, suggesting that metam-sodium or its degradation products are highly volatile. At the 0.1, 1, and 10 mg/animal dose levels, the amounts of metam-sodium absorbed over a 10 hour exposure period were 2.4, 3.7, and 1.5% of the applied dose, respectively. Absorbed radioactivity was either eliminated in urine or exhaled and subsequently trapped in expired air traps. Less than 0.7% of the applied dose was recovered in feces and the recovery of radioactivity from

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the carcass ranged from below the limit of detection to 1.2%. Following applications of metam-sodium at 1 and 10 mg/animal, concentrations of radioactivity in blood and plasma peaked at one hour postdose. Levels of radioactivity in blood and plasma in the 0.1 mg/animal group were below the limit of detection. This study showed that (1) metamsodium is poorly absorbed following a single dermal application; (2) absorbed radioactivity is rapidly excreted, primarily via the urine and expired air; and (3) washing the application site with soap and water effectively removes the majority of the applied dose. A further dermal absorption study with metam-sodium showed that absorption for rats was only 2.5% of the applied dose (U.S. EPA, 1994a). Radiolabeled 14C metam-sodium was applied to shaved dorso-lumbar skin sites of rats at concentrations of 8.6, 86.2, or 862 (g/cm2. Animals were exposed to metam-sodium for 1, 2, 10, 24, or 72 hours. At the end of the study, total dermal absorption after 72 hours was determined to be 2.5% of total applied dose. Since metam-sodium is poorly absorbed dermally, human skin surfaces are acidic, and sweat is approximately pH5, metam-sodium that may come in contact with skin is expected to rapidly degrade prior to absorption.

107.9  Metabolism After oral ingestion, metam-sodium is rapidly absorbed, metabolized, and excreted from the body. Exhalation and excretion in the urine are the major elimination pathways after oral exposure. Metam-sodium is poorly absorbed following dermal application but the metam-sodium that is absorbed is rapidly excreted, primarily through the urine and expired air. In study of biokinetics and metabolism, radiolabeled metam-sodium [14C] (purity  99%) was administered to Sprague–Dawley rats at dose levels of 10 or 100 mg/kg (Hawkins et al., 1987). Blood, urine, and feces were tested for radioactivity for up to seven days postdosing while expired air was collected up to 72 hours postdose. The results of this study showed that metam-sodium was rapidly and completely absorbed after oral ingestion. Radioactivity in plasma reached a maximum level in 1 hour and decreased to near background levels by 24 hours. Animals receiving [14C] metam-sodium at 10 mg/kg eliminated approximately 25% of the total radioactivity through the urine during the first 8 hours. By 168 hours, 55% of the total activity had been eliminated through the urine and 3–4% through feces. At the 100 mg/kg dose, 18% of the total activity had been eliminated in the urine by 8 hours and 40% by 168 hours. Within 24 hours, expired air from 10 mg/kg rats contained approximately 32% of the radioactivity with 1% MITC, 15% carbon disulfide (CS2)/ carbonyl sulfide (COS), and 17% carbon dioxide. At 24 hours for the 100 mg/kg dose level, expired air contained

approximately 48% of the total radioactivity with 24% MITC, 18% CS2/COS, and 6% C02. Negligible amounts of radiolabeled material were expired from 24 to 72 hours at either dose level. Approximately 98% of the radioactivity had been eliminated by the seventh day, with only 2% of the activity remaining in the tissues. The highest concentration of radioactivity was found in the thyroid, but significant concentrations were also found in the liver, kidneys, and lungs. Analysis of the urinary metabolites found that glutathione conjugation with MITC is the source of the major urinary metabolite, N-acetyl-S-(N-methylthiocarbamoyl)-lcysteine, which accounted for 21% of the excreted dose. No evidence for glucuronide or sulfate conjugates of the metabolites of metam-sodium was found. Based on the results of this study, it appears that metam-sodium degrades to either CS2 or MITC in the stomach (accelerated by the stomach pH). MITC is eliminated either through exhalation or in the urine after glutathione conjugation in the liver. CS2 is eliminated by exhalation or further metabolized in the liver to CO2 prior to elimination. Therefore, two different metabolic pathways, CS2 metabolism and MITC conjugation, are involved in urinary elimination. At higher dose levels, saturation of the metabolic processes results in greater exhalation of unmetabolized products. The in vivo dermal absorption study conducted by Stewart (1992), which was described previously, further demonstrated that (1) metam-sodium is poorly absorbed following a single dermal application; (2) absorbed radioactivity is rapidly excreted, primarily via the urine and expired air; and (3) washing the application site with soap and water effectively removes the majority of the applied dose. In both this study and the Hawkins et al. (1987) metabolism study, peak blood and plasma radioactivity levels occurred one hour after dosing.

References Allen (1991). Bomhard, E. (1992). Frequency of spontaneous tumors in Wistar rats in 3-month studies. Exp. Toxic. Pathol. 44, 381–392. Bomhard, E., Karbe, E., and Loesser, E. (1986). Spontaneous tumors of 2000 Wistar TNO/W.70 rats in two-year carcinogenicity studies. J. Environ. Path. Toxicol. Oncol. 7, 35–52. Brammer, A. (1992). “Metam-Sodium: 90-Day Oral Dosing Study in Dogs.” Unpublished study (Rep. CTL/P/3679) conducted by Zeneca Central Toxicology Laboratory, Alderley Park, Macclesfield, Cheshire, UK. Submitted by Metam-sodium Task Force. Brammer, A. (1993). “Metam-Sodium: Assessment of Recovery in Dogs.” Unpublished study (Rep. CTL/L/5204) conducted by Zeneca Central Toxicology Laboratory, Alderley Park, Macclesfield, Cheshire, UK. Submitted by Metam-sodium Task Force. Brammer, A. (1994). “Metam-Sodium: 1-Year Oral Toxicity Study in Dogs.” Unpublished study (Rep. CTL/P/4196) conducted by Zeneca Central Toxicology Laboratory, Alderley Park, Macclesfield, Cheshire, UK. Submitted by Metam-sodium Task Force.

Chapter | 107  Metam-Sodium

Carlock, L. L., Chen, W. L., Gordon, E. B., Killeen, J. C., Manley, A., Meyer, L. S., Mullin, L. S., Pendino, K. J., Percy, A., Sargent, D. E., Seaman, L. R., Svanborg, N. K., Stanton, R. H., Tellone, C. I., and Van Goethem, D. L. (1999). Regulating and assessing risks of cholinesterase-inhibiting pesticides: Divergent approaches and interpretations. J. Toxicol. Environ. Health. B 2, 105–160. Clowes, H. M. (1993). “Metam Sodium: In Vitro Absorption through Rat and Human Skin.” Unpublished study (Rep. CTL/P/4118) conducted by Zeneca Central Toxicology Laboratory, Alderley Park, Macclesfield, Cheshire, UK. Submitted by Metam-sodium Task Force. Crain, R. C. (1958). Spontaneous tumors in the Rochester strain of the Wistar rat. Amer. J. Pathol. 34, 311–335. Deerberg, F., Rapp, K., Rehm, S., and Pitterman, W. (1980). Genetic and environmental influences on lifespan and diseases in Han: Wistar rats. Mech. Ageing Devel. 14, 333–343. Hawkins, D. B., Elsom, L. F., and Girkin, G. (1987). “The Biokinetics and Metabolism of 14C-Metam in the Rat.” Unpublished study conducted by Huntingdon Research Centre, UK. Submitted by BASF Corporation, Research Triangle Park, NC. Hellwig, J. (1987). “Report on the Study of the Prenatal Toxicity of Metam-Sodium (Aqueous Solution) in Rabbits after Oral Administration (Gavage).” Unpublished study (Project 38R0232/8579) conducted by BASF Aktiengesellschaft, Federal Republic of Germany. Submitted by BASF Corporation Chemicals Division, Parsippany, NJ. Hellwig, J., and Hildebrand, B. (1987). “Report on the Study of the Prenatal Toxicity of Metam-Sodium in Rats after Oral Administration (Gavage).” Unpublished study (Rep. 87/0128) conducted by BASF Aktiengesellschaft, West Germany. Submitted by BASF Corporation, Research Triangle Park, NC. Hodge, M. C. E. (1993). “Metam Sodium: Developmental Toxicity Study in the Rabbit.” Unpublished study (Rep. CTL/P/4035) conducted by Zeneca Central Toxicology Laboratory, Alderley Park, Macclesfield, Cheshire, UK. Submitted by Metam-sodium Task Force. Homer, S. A. (1994). “Metam-Sodium: Two Year Drinking Study in Mice.” Unpublished study (Rep. CTL/P/4095) conducted by Zeneca Central Toxicology Laboratory, Cheshire, UK. Submitted by Metamsodium Task Force. JMPR (Joint Meeting on Pesticide Registrations, World Health Organization) (1995). “Pesticide Residues in Food—1995.” FAO Plant Production and Protection Paper 133, p. 4. Jowa, L. (1998). Metam: Animal toxicology and human risk assessment. In “Toxicology and Risk Assessment: Principles, Methods, and Applications” (A. M. Fan and L. W. Chang, eds.), p. 619. Dekker, New York. Klaassen, C. D. (1986). Chapter 2: Principles of toxicology. In “Cassarett and Doull’s Toxicology: The Basic Science of Poisons” (C. D. Klaassen, M. O. Amdur, and J. Doull, eds.), pp. 11–32. McGraw– Hill, New York. Kroes, R., Garbis-Berkvens, J. M., de Vries, T., and van Nesselrooy, H. J. (1981). Histopathological profile of a Wistar rat stock including a survey of the literature. J. Gerontol. 36, 259–279. Lamb, I. C. (1993). “An Acute Neurotoxicity Study of Metam-Sodium in Rats (Definitive).” Unpublished study (Study WIL-188009) conducted by WIL Research Laboratories, Inc., Ashland, OH. Submitted by Metam-sodium Task Force, Los Angeles, CA. Liggett, M. P., and McRae, L. A. (1991). “Skin Irritation to Rabbits with Metam-Sodium.” Unpublished study (Study 90997D/UCB 368/SE) conducted by Huntingdon Research Centre, Ltd., UK. Submitted by UCB Chemicals Corporation, Norfolk, VA.

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Mackay (1996). Merck (1989). Metham sodium. In “The Merck Index” (S. Budavari, M. L. O’Neil, A. Smith, and P. E. Heckelman, eds.) 11th ed., p. 937. Merck, Rahway, NJ. Milburn, G. M. (1993). “Metam Sodium: Multigeneration Study in the Rat.” Unpublished study (Rep. CTL/P/3788) conducted by Zeneca Central Toxicology Laboratory, Alderley Park, Macclesfield, Cheshire, UK. Padgett, E. L., Barnes, D. B., and Pruett, S. B. (1992). Disparate effects of representative dithiocarbamates on selected immunological parameters in vivo and cell survival in vitro in female B6C3F1 mice. J. Toxicol. Environ. Health 37, 559–571. Parcell, B. I., and Denton, S. M. (1991). “Delayed Contact Hypersensitivity in the Albino Guinea Pig.” Unpublished report (Rep. 901002D/UCB370/SS) conducted by Huntingdon Research Centre, Ltd., UK. Submitted by UCB Chemicals Corporation, Norfolk, VA. Pruett, S. B., Barnes, D. B., Han, Y. C., and Munson, A. E. (1992). Immuno-toxicological characteristics of sodium methyldithiocarbamate. Fund. Appl. Toxicol. 18, 40–47. Rattray, N. J. (1994). “Metam-Sodium: Two Year Drinking Study in Rats.” Unpublished study (Project PR0838) conducted by Zeneca Central Toxicology Laboratory, Cheshire, UK. Submitted by Metamsodium Task Force. Rehm, S., Deerberg, E., and Rapp, K. G. (1984). A comparison of lifespan and spontaneous tumor incidence of male and female Han: WIST virgin and retired breeder rats. Lab. Anim. Sci. 34, 458–464. Stewart, F. P. (1992). “Metam-Sodium: In Vivo Percutaneous Absorption Study in the Rat.” Unpublished study (Report 7268-38/142) conducted by Hazleton UK, Harrogate, North Yorkshire, UK. Thomassen, R. W. (1998). “Review of the 2-Year Drinking Water Study with Metam-Sodium in the Wistar Rat.” Unpublished report submitted to the California Office of Environmental Health Hazard Assessment. Reviewed with California Office of Environmental Health Hazard Assessment at a conference on July 29, 1998. Tinston, D. J. (1993). “Metam Sodium: Developmental Toxicity Study in the Rat.” Unpublished study (Report CTL/P/4052) conducted by Zeneca Central Toxicology Laboratory, Alderley Park, Macclesfield, Cheshire, UK. United States Environmental Protection Agency (U.S. EPA). (1991). “Metam-Sodium—Review of Two Developmental Toxicity Studies in Rats and Rabbits Submitted by the Registrant.” Memorandum from Y M. Ioannou to S. Lewis. United States Environmental Protection Agency (U.S. EPA). (1992). “Metam-Sodium—Review of a 90-Day Study in Mice.” Memorandum from Y M. Ioannou to S. Lewis. United States Environmental Protection Agency (U.S. EPA). (1993). “Sodium N-Methyldithiocarbamate (Metam-Sodium).” Memorandum from T. F. McMahon to A. Mehta. United States Environmental Protection Agency (U.S. EPA). (1994a). “Worker and Residential/Bystander Risk Assessment of Metamsodium During Soil Applications.” Memorandum from A. Mehta to J. Ellenberger and J. House-nger. United States Environmental Protection Agency (U.S. EPA). (1994b). “Metam-Sodium: Review of a Chronic Toxicity/Carcinogenicity Study in Rats and Chronic Toxicity Study in Dogs Submitted by the Registrant.” Memorandum from T. F. McMahon to T. Myers. United States Environmental Protection Agency (U.S. EPA). (1994c). “Metam-Sodium: Review of a Mouse Carcinogenicity Study Submitted under FIFRA Section 6(a)(2) by the Registrant.” Memorandum from T. F. McMahon to T. Myers.

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United States Environmental Protection Agency (U.S. EPA). (1994d). “Metam-Sodium: Review of an Acute Neurotoxicity Study Submitted by the Registrant.” Memorandum from T. F. McMahon to T. Myers. United States Environmental Protection Agency (U.S. EPA). (1994e). “Metam-Sodium: Review of a Subchronic Neurotoxicity Study in Rats.” Memorandum from T. F. McMahon to T. Myers. United States Environmental Protection Agency (U.S. EPA). (1995). “Carcinogenicity Peer Review of Metam-Sodium.” Memorandum from T. F. McMahon and E. Rinde to L. Cole and T. Myers. United States Environmental Protection Agency (U.S. EPA). (1997). “Case Study—Methyl Bromide Alternative: Metam-Sodium as an

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Alternative to Methyl Bromide for Fruit and Vegetable Production.” US EPA document posted at http://earthl.epa.gov/ozone/mbr/ metams.htm. Whiles, A. J. (1991). “Metam-Sodium: 90-Day Drinking Water Study in Mice with a 28-Day Interim Kill.” Unpublished Report (Rep. CTL/ P/3185) conducted at ICI Central Toxicology Laboratory, Alderley Park, Macclesfield, Cheshire, UK. World Health Organization (WHO) (1990). “Environmental Health Criteria 104. Principles for the Toxicological Assessment of Pesticide Residues in Food.” International Program on Chemical Safety, World Health Organization, Geneva.

Chapter 108

Methyl Iodide D.J. Ashworth and S.R. Yates USDA-ARS United States Salinity Laboratory, Riverside, California

108.1  Background Methyl iodide (MeI, iodomethane, CH3I) was reported as a potential alternative to the stratospheric ozone-depleting fumigant methyl bromide (MeBr) in the mid-1990s (Sims et al., 1995; Ohr et al., 1996). It has since received significant research attention to determine its environmental fate and transport as well as its pesticidal efficacy. As a preplant soil fumigant, MeI can be used alone or in combination with chloropicrin to control plant pathogens, nema­ todes, insects, and weeds on crops such as strawberries, tomatoes, peppers, ornamentals, turf, trees, and vines (U.S. EPA, 2008). In addition to offering a similar pesticidal efficacy to MeBr, MeI also offers the major advantage of having the very short atmospheric half-life of 4–8 days, compared to 1.5–2.0 years for MeBr (Gan et al., 1997), due to rapid photolysis when exposed to UV radiation (Gan and Yates, 1996). Importantly therefore, it is not considered to degrade stratospheric ozone and so its emissions from soil are of less concern compared to methyl bromide. Nevertheless, MeI fumigation has the potential to increase human health risks through direct inhalation of the MeI gas and due to its potential role in the formation of nearsurface ozone. Ruzo (2006) listed several advantages of MeI over MeBr for use as an agricultural fumigant as: Not an ozone depletion chemical, nor a potential ground water problem l Liquid form, easier to handle, decreased worker exposure l Readily biodegradable l Equal or greater toxicity to target organisms on a molar basis

Synergy with chloropicrin is comparable to methyl bromide l Comparable spectrum of activity; therefore, use as a ‘cocktail’ not envisioned beyond combinations with chloropicrin comparable to those of methyl bromide l

In this chapter, the properties, registration/health issues, efficacy and soil fate and transport of MeI will be considered. Since MeBr is often considered the standard by which alternative fumigants are assessed, comparisons between MeI and MeBr are frequently drawn.

108.2  Methyl iodide properties Methyl iodide is a clear, colorless liquid with an acrid odor. It has the chemical structure shown in Figure 108.1. Selected properties of MeI, together with those of several other commonly used soil fumigants, are shown in Table 108.1. Compared to the other MeBr alternatives listed in the table, it is evident that the properties of MeI most closely match those of MeBr, suggesting it may be the most promising alternative to MeBr for soil fumigation. Indeed, in reviewing the status of chemical alternatives to methyl bromide for preplant fumigation of soil, Duniway (2002) concluded that H

l

Hayes’ Handbook of Pesticide Toxicology Copyright © 2010 Elsevier Inc. All rights reserved

H

C

I

H Figure 108.1  Chemical structure of methyl iodide.

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Table 108.1  Selected Properties of Important Soil Fumigantsa Water solubility at 20°C (% wt/wt)

Vapor pressure at 20°C (mmHg)

Boiling point (°C)

Henry’s constant (KH) (air/water)

Half-life in soil (days)

MeI

1.40

400

42

0.210

20

MeBr

1.34

1420

4

0.244

22

PrBr

1.49

72

88

0.046

5

MITC

0.76

21

119

0.011

7

1,3D

0.22

34

104

0.056

11

CP

0.20

18

112

0.093

1

a

Data from Ajwa et al., 2002.

MeI and propargyl bromide (Yates and Gan, 1998) probably have activities that most closely parallel that of MeBr in soil. The relatively high Henry’s constant and vapor pressure of MeI indicates that it is likely to persist in the gaseous phase and readily diffuse through porous media. Additionally, its long half life in soil ensures that long contact times with soil pests can be achieved. These properties are indicative of effective fumigant pesticides.

Table 108.2  Size of Buffer Zones for Varying MeI Application Scenariosa Application rate (lb a.i./treated acre)

Size of contiguously treated area (acres)

Buffer zone distance (ft) if HDPE or LDPE used

175 (max)

20 to 40

500

10 to 20

300

5 to 10

100

Up to 5

50

20 to 40

375

10 to 20

225

5 to 10

75

Up to 5

40

20 to 40

250

10 to 20

150

5 to 10

50

Up to 5

25

20 to 40

125

10 to 20

75

5 to 10

25

Up to 5

25

108.3  Registration issues As the product Midas (Arysta LifeScience, NC, USA), MeI is commercially available in formulations with chloro­ picrin (CP) at ratios of 98:2, 50:50, 33:67, and 25:75 (MeI: CP). In September 2006, an experimental use permit was granted for MeI by the United States Environmental Protection Agency (U.S. EPA). This permit was for limited use on 1000 total acres in Florida, Georgia, Michigan, North Carolina, South Carolina, Tennessee, and Virginia. In October 2007, the U.S. EPA approved the 1-year registration of MeI as a soil fumigant, and in July 2008, Florida became the 45th U.S. state to approve MeI for commercial use. In October 2008, the U.S. EPA extended the registration of MeI. No time limit was established to allow for coordination of this registration with implementation of the reregistration decisions for other soil fumigants (U.S. EPA, 2008). Currently, MeI is registered for use in 47 states. (Arysta Lifesciences, 2008). In California, the Department of Pesticide Regulations (DPR) took the decision to carry out its own risk assessment of MeI use and is due to report the outcome of this process at the end of 2009. No decision on the registration of MeI within California will be taken before this risk assessment is complete (California DPR, personal communication). A number of restrictions accompanied the U.S. EPA registration to mitigate the risk associated with MeI use. Most notable of these restrictions is the requirement for the establishment of buffer zones around treated areas. The

131 (75%)

88 (50%)

44 (25%)

a

Data from USEPA, 2008. HDPE: high-density polyethylene. LDPE: low-density polyethylene.

size of such buffer zones is related to the acreage treated and the amount of MeI applied (Table 108.2). As seen in Table 108.2, the maximum allowable application rate is 175 lb a.i./treated acre. Other notable restrictions are that use within ¼ mile of any occupied sensitive site such as a

Chapter | 108  Methyl Iodide

school, day care facility, nursing home, hospital, prison, or playground is prohibited, and that entry to the treated area is restricted for a period of 5 days. Despite the USEPA restrictions on the use of MeI, concerns within the wider scientific community over the potential health risks of commercial MeI use have been noted. In a letter to the U.S. EPA dated September 24, 2007, Dr R.G. Bergman (University of California, Berkeley) called for a prevention of the registration of MeI as a pesticide on the basis that release of MeI from fumigated soils would result in human exposure and increased health risk. The letter pointed out the cancer hazard of alkylating agents like MeI, and also the potential for thyroid toxicity, permanent neurological damage, and fetal loss in experimental animals. In addition, the letter questioned the risk assessment procedures employed by the U.S. EPA. The letter was co-signed by a further 53 scientists and documented to represent personal, rather than institutional, opinions. In a response dated October 5, 2007, the U.S. EPA noted that their analysis of MeI over the previous 4 years was “one of the most thorough analyses ever completed on a new pesticide.” In relation to potential health risks, the U.S. EPA reported that in their review, the only evidence of carcinogenicity following exposure to MeI was related to thyroid cancer. However, their risk assessments indicated that exposures expected from MeI use would be well below those that would cause thyroid effects leading to cancer. Similarly, the most sensitive endpoints (nasal irritation, fetal loss, and neurotoxicity) were not considered to be of concern at the expected levels of MeI exposure.

108.4  Exposure issues Methyl iodide is a highly toxic chemical with the potential to cause serious health effects. The Material Safety Data Sheet (MSDS) for MeI gives the toxicity data: oral-rat LD50 of 76 mg/kg; intraperitoneal-rat LD50 of 101 mg/kg; inhalationrat LC50 of 1300 mg/m/3/4 h; subcutaneous-mouse LD50 of 110 mg/kg. The MSDS also states that health effects associated with the chemical are potentially severe (cancer-causing). It has been shown to cause cancer in animals and may be linked to cancer in humans as well as damage to the central nervous system. Health effects can be caused by inhalation, ingestion, skin contact, and eye contact. In relation to human exposure to inhalants, the American Conference of Governmental Industrial Hygienists (ACGIH) report threshold limit values (TLV). Of these, the time-weighted average (TWA) is the level at which an “average” worker can be safely exposed repeatedly (i.e., 8 h/day, 40 h/week). The TLV-TWA for MeI is 2 mg/l (ACGIH, 1994). This compares to other, more commonly encountered inhalants thus: bleach (sodium hypochlorite), 1 mg/l (as Cl2); ammonia, 25 mg/l; ozone, 0.05 mg/l (for heavy work) to 0.1 mg/l (for light work). On this basis, MeI would appear less hazardous than

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both bleach and ozone. Nevertheless, it is important that MeI fumigation practice be managed in such a way as to limit its potential risk to humans, particularly agricultural workers and nearby populations. Potential human health effects would be associated primarily with inhalation. According to the MeI MSDS, inhalation may irritate the respiratory tract and overexposure may produce symptoms of vertigo, delirium, and mental disturbances. Other symptoms may include slurred speech, drowsiness, coughing, nausea, and vomiting. Higher exposures can cause a build-up of fluid in the lungs, which can cause death.

108.4.1  Quantifying Exposure of   Plant Pests to MeI The diffusion of soil fumigants has been studied quantitatively since the 1950s. For example, Hemwall (1959) was one of the first to utilize computers to numerically solve an equation to simulate and predict fumigant diffusion in soil. This allowed the theoretical study of both the diffusion pattern of the chemical and also allowed the calculation of a biological control function, which is commonly called the concentration-time index: t0

CT (t 0 ) 

∫ C x ,y dt 0

(Wang and Yates, 1999)

where CT(t0) is the concentration time index (g/h/cm3) up to time t0 (h); Cx,y is the chemical concentration at a given location (x,y) in the soil air (g/cm3); and t is time (h). The value of CT can be related to the level of biological control, i.e., organism survival under the conditions studied. Such simulations can be used to obtain information on the effectiveness of application methods while ensuring that the soil fumigant distribution is not compromised and that efficacious dosages and biological control would be achieved.

108.4.1.1  Weed Control Zhang et al. (1998) found that the optimal soil moisture for MeI to kill the weed species, Abutilon theophrasti and Lolium multiflorum, in sandy soils was 14% water content, and that greater efficacy was obtained when the temperature during fumigation was above 20°C. The same authors also noted that 100% mortality of weeds was achieved more rapidly with MeI (24 h), compared to MeBr (36 h). Across a range of environmental factors, these authors concluded that on a molar basis, MeI was consistently more effective than MeBr. Similarly, Ohr et al. (1996) and Zhang et al. (1997) concluded that MeI appears to be a suitable replacement for MeBr because it can be used in similar situations and has a superior efficacy against a broad spectrum of weed pests. In the control of yellow

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nutsedge tubers, Hutchinson et al. (2004) reported that MeI was around 2.6 times more potent that MeBr, and that combining MeI with 17% chloropicrin resulted in a synergistic interaction and an improved control of the pest. The same authors also noted that MeI produced similar control of the pest to a combination of 1,3-dichloropropene with 17% chloropicrin (essentially the commercial product, Telone C-17). Unruh et al. (2002) found that MeI controlled weedy grass species, sedge species, and broadleaf weeds present at two locations under differing environmental conditions. However, Webster et al. (2001) noted that while MeI was effective in controlling purple nutsedge early in the growing season of a bell pepper–squash cropping sequence, by the end of the growing season, the MeI offered no advantage over a control. Experiments performed in our laboratory studied the survival of barnyard grass seed subjected to differing concentrations of MeI in soil. To 50-g samples of moist (12% v/v) sandy loam soil (Buttonwillow, CA) each containing the same quantity of barnyard grass seed, MeI was added to give soil concentrations ranging from 0 to 41 mg/kg. After a period of 24 h, the survival of the seeds was determined. Results are shown in Figure 108.2, where the CT value vs. barnyard grass seed survival is plotted. The lethal dose at which 50% of the seed population was killed (LD50) was 268 g/h/cm3.

of plant-parasitic nematodes throughout the first growing season similar to the control achieved with methyl bromide (at 507 kg/ha). Eayre et al. (2000) also reported that very similar reductions in population densities of the nematode Paratylenchus were observed for both MeI and MeBr fumigation. Becker et al. (1998) evaluated methyl iodide as a soil fumigant in container and small field plot studies. They concluded that MeI was an excellent soil fumigant. Compared to MeBr, their container studies showed that MeI was significantly more effective against several species of plant parasitic nematodes at equal molar rates. Moreover, in their field plots, soil populations of root-knot nematodes were no longer detected after methyl iodide fumigation at application rates of 112 kg/ha and above. Nevertheless, only at their higher application rates (168 kg/ ha and above) was significant root-knot galling prevented in lima bean grown for 2 months after fumigation. Using the approach described in the previous section for barnyard grass seed, experiments in our laboratory have also studied the survival of citrus root nematodes at varying MeI concentrations in soil. For MeI concentrations ranging from 0 to 6 mg/kg, the CT value was plotted against nematode survival (Figure 108.3), and it was determined that the LD50 for the nematodes was 19 g/h/cm3. This indicates that MeI is more efficacious toward nematodes than to the barnyard grass seeds.

108.4.1.2  Nematode Control

108.4.1.3  Pathogenic Microbe Control

In relation to nematode control, Webster et al. (2001) found that nematode-susceptible pepper and squash plants under MeI treated conditions exhibited lower root-gall indices (root knot nematode) than the control. Compared to MeBr, Hutchinson et al. (1999a) found that MeI was 4.67 and 1.77 times more potent than MeBr in laboratory and field conditions, respectively. Schneider et al. (2006) found that in vineyard soils, a 50:50 combination of MeI and chloropicrin drip-applied at 269 kg/ha provided control

For field plots located in main strawberry production areas of California, Stromberger et al. (2005) reported that, in common with other soil fumigants (including MeBr), MeI eliminated soil-borne fungal pathogens and reduced culturable fungal populations up to 4 weeks after fumigation. Soil microbial respiration, enzyme activity, and potential nitrification rates were also decreased with fumigant application, indicating a significant impact of the fumigants on

Figure 108.2  Survival of barnyard grass in sandy loam soil after a 24-h MeI fumigation at 20°C.

Figure 108.3  Survival of Tylenchulus semipenetrans (citrus nematode) in sandy loam soil after a 24 h MeI fumigation at 20°C.

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the microbial flora and fauna. Even though Webster et al. (2001) found that the number of total fungi isolated from soil was not reduced by MeI fumigation, the laboratory studies of Hutchinson et al. (2000) suggested that MeI was, on average, 2.7 times more efficacious than MeBr at killing fungal species. Furthermore, the same workers found that, in combination with chloropicrin, MeI was 2.8 times more efficacious against the fungus Fusarium oxy­ sporum than MeI alone. Stanghellini et al. (2003) injected MeI to melon beds as a hot gas through drip irrigation tape (448 kg/ha) and observed equal or better reductions in the fungal pathogen, Monosporascus cannonballus, when compared to MeBr applied at the same rate. Similarly, Ohr et al. (1996) found that for controlling the plant fungi Phytophthora citricola, P. cinnamomi, P. parasitica, and Rhizoctonia solani, MeI was equal to, or better than, MeBr. Becker et al. (1998) also found that the fungal species R. solani was more strongly controlled by MeI than MeBr. For fumigant rates of 88.6, 177.2, and 354.4 mmol/m2, these workers found R. solani survival of 92.5, 77.5, and 10.0%, respectively, for MeBr, and 27.5, 7.5, and 0%, respectively, for MeI. Ibekwe et al. (2007) reported that MeI (and MeBr) were effective in reducing both the concentration of Escherichia coli 0157 in soil, and the survival of the pathogen on lettuce leaf surface, suggesting that the fumigants may have played some role in reducing the transfer of E. coli 0157 from soil to leaf. As previously described for barnyard grass and citrus nematodes, our laboratory has studied the effects of MeI on fungi in soil. MeI concentrations ranging from 0 to 111 mg/ kg produced the Fusarium oxysporum mortality curve (CT value vs. survival) shown in Figure 108.4. The high LD50 value (1206 g/h/cm3) suggests that MeI is less effective in the control of this organism than to both the barnyard grass and, particularly, the citrus nematodes. These results indicate that relatively large application rates may be required for fungi control. Alternatively, the combination of MeI

Methyl Iodide Fusarium oxysporum

100

Survival (%)

80 60 40 20 0 0

500

1206

1000

1500

Concentration-Time

2000

2500

3000

(µg/h/cm3)

Figure 108.4  Survival of Fusarium oxysporum (fungi) in sandy loam soil after a 24-h MeI fumigation at 20°C.

with a fumigant known to offer improved fungi control (e.g., chloropicrin) may be required.

108.5  Crop production By controlling pests, fumigants are expected to exert a beneficial effect upon the growth and production of crop plants. For example, Hutchinson et al. (1999b) reported that plots fumigated with either MeBr or MeI produced at least 161 and 181% more marketable carrots without nematode damage, respectively, than plants in control plots. Also for strawberries, Kabir et al. (2005) found that runner plant production was increased by both MeBr/chloropicrin and MeI/chloropicrin soil fumigation, compared to a control. Moreover, fruit yields from plants grown on MeI/ chloropicrin-fumigated soils were similar to those grown on the MeBr/chloropicrin soils. In peach replant soils fumigated with MeBr and MeI, Eayre et al. (2000) found that both fumigants resulted in increased tree trunk diameter and branch weight over control levels. Moreover, plots fumigated with MeI did not differ from plots fumigated with MeBr in trunk growth or weight of branch prunings.

108.6  Environmental fate of methyl iodide as a soil fumigant 108.6.1  Transformation in Aqueous Solution Compared to its half-life in water (74 days), Gan and Yates (1996) observed a marked reduction in half-life for MeI when incubated in aniline solution (4.9 days). The more rapid reaction with this model nucleophilic compound suggested that degradation of the MeI occurred via nucleophilic substitution. The end products N-methylaniline (84%), N,N dimethylaniline (13%), and residual MeI (3%) were identified by gas chromatography-mass spectrometry (GC-MS) after 25 days of incubation. Similarly, Zheng et al. (2003) noted that MeI half-life in water was reduced from 108 days in the presence of 1.0 mM concentrations of a range of agricultural fertilizers and nitrification inhibitors. Most significant reductions were observed for sodium diethyldithiocarbamate and ammonium diethyldithiocarbamate, with half-lives of 0.13 and 0.14 days, respectively. The authors also found that, as the initial concentration of these chemicals increased, the first-order degradation rate constant of MeI also increased. Indeed, a positive, linear (r2  0.99) relationship was reported. The authors propose that in aqueous solutions of these chemicals, (C2H5)2NSCS exists as the strongly nucleophilic thiolate ion, which is susceptible to SN2 nucleophilic substitution reactions with halogenated hydrocarbons, e.g.: CH3I  (C2 H 5 )2 NSCS   → (C2 H 5 )2 NSCS-CH3  I

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Zheng et al. (2004) studied the effect of thiourea (a nitrification inhibitor) on MeI transformation in aqueous solution at various pH values. These authors found that both in the presence and absence of thiourea, neutral and base-catalyzed hydrolysis of MeI significantly decreased the half-life of the fumigant as pH increased above circum-neutrality (from 381 h at pH 6.9 to 106 h at pH 12). It was suggested that this was due to the relatively high nucleophilic reactivity of OH. Additionally, compared to a control (no thiourea), MeI dissipation was markedly enhanced in the presence of thiourea (1.0 mM) at a range of pH values (4–12). For example, at pH 4.0, approximately 70% of the MeI had disappeared after 5 days in the presence of thiourea, compared to 20% in the control. The enhanced dissipation of MeI in the presence of thiourea was attributed to a nucleophilic substitution reaction with thiourea. However, it appeared that the reaction between MeI and thiourea was pH-independent and that a change in pH might not significantly affect the MeI transformation process in aqueous systems. Bondarenko et al. (2006) studied the dehalogenation of halogenated fumigants by polysulfide salts. They observed that when polysulfide concentration in aqueous solution was increased from 0.2 mM to 1.0 mM, the disappearance time values for 0.2 mM MeI decreased from 27.2 to 2.2 h. The authors suggested that the reaction here is also SN2 nucleophilic substitution and highlight the potential for use of polysulfide salts as a pollution mitigation strategy such as for disposal of fumigant wastes, treatment of fumigantcontaining wastewater, and cleanup of fumigant residues in environmental media. In studying aqueous clay suspensions at pH 6.9, Zheng et al. (2004) concluded that sorption onto minerals such as montmorillonite and kaolinite did not lead to a more rapid dissipation of MeI.

108.6.2  Transformation in Soil Since it is an irreversible process, degradation of MeI in soil is an important factor in determining the fraction of the fumigant available for transport. Gan and Yates (1996) studied the degradation of MeI in three California soils and a nursery potting mix under sterile (autoclaved) and nonsterile conditions. For Carsetas loamy sand, Linne clay loam, a potting mix, and Greenfield sandy loam, MeI halflives were reported as 11, 13, 13, and 43 days, respectively, under nonsterile conditions. In each soil, these values were approximately twice those obtained for MeBr in parallel experiments. Upon soil sterilization, MeI half-life in each soil was 10, 9, 17, and 63 days, respectively. These authors concluded that half-lives were not drastically affected by soil sterilization. The same conclusion was drawn by Zheng et al. (2004), suggesting that MeI is degraded primarily by chemical, rather than biological, processes. Such processes may be facilitated by the presence of soil organic matter and, more specifically, the fumigant-degrading nature of functional

Hayes’ Handbook of Pesticide Toxicology

groups present in the molecular structure of the organic matter. For example, in the work of Gan and Yates (1996), those soils relatively rich in organic matter (Linne and Carsetas; 2.99 and 2.51% organic matter, respectively) yielded halflives of 9–13 days, compared to 43–63 days in the lower organic matter Greenfield soil (0.92% organic matter). Nevertheless, the quality of the organic material may have been more significant than its quantity in determining MeI degradation, since the potting mix which had by far the highest organic matter (9.60%) did not produce the shortest halflife values (13–17 days). The authors therefore suggested that the nondegraded plant residues of the potting mix may not have been very effective in inducing MeI degradation. Zheng et al. (2003) examined the effects of a range of agricultural fertilizers (oxamide, calcium cyanamide, and urea), and nitrification inhibitors (dicyandiamide, thiourea, 1-allyl-2-thiourea, ammonium thiosulfate, sodium diethyldithiocarbamate, and ammonium diethyldithiocarbamate) on the degradation of MeI in soil. Compared to control soil (half-life of 309 h), they observed that MeI half-life was reduced to between 16 and 227 h across all treatments. The most marked reductions in half-life were found for the chemicals ammonium thiosulfate (ATS), thiourea, and 1-allyl-2-thiourea (17, 16, and 16 h, respectively). The authors attributed the effectiveness of ammonium thiosulfate to its existence (as the thiosulfate ion) in the aqueous soil phase where reaction with the fumigant takes place. ATS is unstable in soil, because the thiosulfate ion can be easily oxidized to sulfate (which is not reactive with MeI); therefore, Zheng et al. (2003) suggested that in order to impact emissions from soil, ATS application should be carried out concurrently with fumigant application. The dissipation half-life of MeI in a range of soils and sand was reported to be between approximately 40 h in Florida muck (organic carbon content 45.8%) to 600 h in sand (presumably 0% organic carbon) (Zheng et al., 2004). This large difference in half-life suggests that organic matter played an important role in degrading the MeI in the Florida muck. Two sandy loam soils yielded intermediate dissipation half-lives of around 350 h and a fine sand of around 430 h. When thiourea was added to these systems at a 2:1 ratio (thiourea: MeI), the dissipation half-lives decreased by factors of 12 (sand), 13 (fine sand), 14–17 (sandy loams), and 5 (muck). These data indicate that the thiourea exerted a strong influence on the rate of MeI transformation. Zheng et al. (2004) therefore suggested that treating MeI fumigated soil with thiourea-amended muck soil may be a useful approach to reducing fumigant emissions.

108.6.2.1  Effects of Temperature and Moisture Content on Transformation in Soil Since soil temperature and moisture content are likely to vary considerably under field conditions (e.g., in response to diurnal variation in solar energy), the effect of these

Chapter | 108  Methyl Iodide

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factors on the transformation of MeI has been studied. In general, the fumigant degradation rate increases at higher temperature and moisture content, presumably due to increased chemical reactivity and microbial activity under such conditions. Indeed, in laboratory experiments, Zheng et al. (2004) found that MeI dissipation half-life decreased in response to increasing temperature, with half-lives of 41.8, 31.4, 18.9, 12.1, and 5.5 h recorded at 4, 10, 20, 30 and 40°C, respectively. Since MeI degradation is not thought to be strongly controlled by microbial processes (Gan and Yates, 1996; Zheng et al., 2004), it is likely that the higher temperatures increased the rate of chemical reactions between the soil and MeI (i.e., Arrhenius-type behavior). In relation to the effect of soil moisture content on pesticide degradation, Walker et al. (1992) and Parkin and Shelton (1994) found a positive correlation. However, in the presence of thiourea, Zheng et al. (2004) found that MeI degradation decreased with increasing soil moisture content. MeI half-lives were 6.5, 9.7, 12.3, 21.3, and 23.6 h, respectively, when thiourea-amended soil moisture contents were 2, 5, 8, 12, and 16%. They suggested that a soil catalytic reaction mechanism of MeI with thiourea on the surface of soil particles was responsible, i.e., not soil moisture stimulating microbial growth.

108.6.2.2  Mechanism of Transformation in Soil Transformation of MeI has been shown to follow first-order kinetics (Gan and Yates, 1996; Zheng et al., 2003, 2004). According to Gan and Yates (1996), MeI may undergo bimolecular nucleophilic substitution (SN2) with H2O.

CH3 I  H 2 O   → CH3 OH  I  H



The same authors noted that degradation of MeI in soils rich in organic matter was considerably faster than the reported 50- to 110-day half-life for MeI hydrolysis in water. Zheng et al. (2004) found that MeI dissipation half-life was 15 times lower in highly organic soil (45.8% organic carbon) compared to sand. Reactions of MeI with nucleophilic functional groups (e.g., -NH2, -NH, -SH, -OH) on soil organic matter were therefore suggested to strongly influence the degradation of MeI in soil (i.e., in a similar fashion to that established for MeBr):

CH3 I  OM-NH 2   → OM-NH-CH3  I  H

In general, MeI has been shown to exhibit a relatively long half-life in soil and a range of aqueous solutions. The persistence of MeI in soils may therefore be expected to lead to the loss of the fumigant from the soil body, i.e., via gaseous phase loss of MeI from the soil surface to the atmosphere, and liquid phase leaching of MeI to groundwater. Such loss pathways are likely to in turn lead to an increase in the environmental risk associated with MeI use.

108.6.2.3  Emissions of MeI from Soil No published work was found that has determined the emission behavior of MeI under field conditions, representing a significant gap in our knowledge. However, investigation of MeI emissions from soil has been carried out using sealed laboratory soil columns. Generally, in such studies, the fumigant is injected at a 30- or 46-cm (12- or 18-in.) depth into soil packed into a stainless steel column. This approach simulates a shank, or broadcast, fumigant injection. An emissions chamber is sealed onto the surface of the column and clean air swept through this chamber, across the soil surface. Emitted MeI is thus pulled with the air flow onto adsorbent charcoal tubes which retain the fumigant for quantitative analysis. A low airflow is maintained so as not to create a negative pressure across the soil surface that might serve to increase the emission of fumigant due to an upward convective flux. In a sealed system such as this, only surface emissions and degradation account for fumigant loss. Charcoal tubes are generally extracted with a solvent such as acetone and the extract analyzed by gas chromatography (GC) using a micro-electron capture detector (ECD). A laboratory system for studying both the emissions and soil distribution of MeI is shown in Figure 108.5. An example of the data produced using such an approach is shown in Figure 108.6, where compared are the MeI emission fluxes from bare soil and high-density polyethylene (HDPE)-tarped soil over time. Tarping was observed to reduce the peak MeI emission by around 50% during the period immediately following fumigant application. Although tarping was observed to reduce total emissions of MeI, it is also noticeable that the emission fluxes over the latter part of the experiments were greater in the tarped treatment. In the published literature, Gan et al. (1997) reported a series of column experiments to compare soil emissions of MeI and MeBr under differing soil types and surface tarps. They observed that total MeI loss from the soil was generally greater than MeBr under the same conditions. For a sandy loam soil, these authors reported 94, 90, and 75% total emissions (expressed as a percentage of the total added to the system) of MeI under control, polyethylene tarped and virtually impermeable film (Hytibar) tarped conditions, respectively. These can be compared to values of 75, 68, and 45%, respectively, for MeBr. For both fumigants, these values are high in relation to emission values for other fumigants under comparable conditions. For example, Hytibar film has been shown to reduce emissions of the fumigant chloropicrin from 82% in a control to just 4% (Gan et al., 2000). In a loamy sand, clay loam and nursery potting mix under polyethylene tarp, emissions of 38, 53, and 33%, respectively, were observed for MeI (Gan et al., 1997). Again, these were higher than the values for MeBr of 32, 41, and 30%, respectively. Differences were attributed to the likely slower soil degradation (longer half-life) of the MeI than MeBr, which

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Emission chamber Air inlet

Stainless steel column

Teflon tubing

Soil

150 cm

Soil gas sampling ports

Manifold Solenoid valves Charcoal tubes Manifold

Datalogger 21X

Flow-meter Vacuum (~150 ml/min)

12 cm Figure 108.5  Laboratory design for studying MeI behavior in soil.

MeI volatilization flux (µg/m2/s)

100 80 Bare soil HDPE

60 40 20 0 0

2

4

6 8 10 Time (days)

12

14

16

Figure 108.6  Time-wise trends in MeI emission fluxes from bare and tarped soil columns after fumigation.

left more MeI for emissions. In terms of emission flux rates, these workers found similar values for both MeI and MeBr. They reported peak fluxes of around 4.5 mg/h/column under nontarped conditions, 2 mg/h/column for polyethylene tarp, and around 0.8 mg/h/column for Hytibar tarp. Notably, the emission peaks were delayed under tarped conditions, particularly for the Hytibar treatment, due to the retardation of fumigant diffusion across the tarp boundary. Comparing the differing soil types they studied under polyethylene tarp, these authors observed that loamy sand and clay loam gave

peak fluxes of around 2 mg/h/column, compared to around 1 mg/h/column for potting mix. The lower values for the potting mix were likely associated with the greater degree of fumigant degradation associated with its greater organic matter content (9.6%) compared to the loamy sand (2.51%) and clay loam (2.99%). Clearly, although the high vapor pressures and Henry’s constants of MeI and MeBr lead to a high degree of pesticidal efficacy, these properties also result in high emissions from soil when compared to other alternative fumigants. Although emissions of MeI are not significant in relation to stratospheric ozone depletion, high emissions are likely to be of concern in relation to worker and local population exposure and the potential for associated health risks. In addition, high levels of emissions may also be an issue in relation to the release of volatile organic compounds, which can be significant in the formation of near-surface ozone, and be the source of a respiratory human-health risk.

108.6.2.4  Soil Distribution The column system described above for determining the emissions of fumigants from soil also allows for the removal of soil gas samples from various depths to determine the time-wise distribution of the fumigant within the soil (Figure 108.4). Extracted soil gas samples can be injected into empty vials for analysis using GC-Headspace equipment, or injected into vials containing organic solvent

Chapter | 108  Methyl Iodide

and analyzed as for the charcoal tube extracts (described above). In a field setting, sampling tubes inserted at various soil depths within a fumigated plot or field can be used for the same purpose. Using a syringe attached to these tubes, a known volume of soil gas can be pulled through charcoal tubes that adsorb MeI. Charcoal tubes are then extracted and analyzed as described above. An example of soil MeI distribution data derived using the soil column approach is shown in Figure 108.7. Here, a bare soil and HDPE tarped soil are compared over time for a 30-cm-depth injection of MeI. The fumigant is seen to rapidly diffuse throughout the soil column from the point of injection. With a tarped soil, the gas becomes trapped below the tarp, leading to relatively high concentrations near the soil surface; such an effect is not seen for the bare soil due to loss via soil-to-air emissions. These data suggest that tarping is necessary to maintain MeI in the soil. This offers the benefits of both reduced soil-to-air emissions, and improved soil pest control close to the soil surface. Using a column approach, Guo et al. (2004) determined the soil distribution of MeI under VIF-tarped conditions for both drip and shank applications at 20 cm depth. They observed that maximum MeI concentration in the soil air was present at the depth of injection through a 168-h period. In general, the shank injection led to greater concentrations and a maximum was reached at 2 h (around 47 mg/l). In the drip application, the maximum was reached at 4 h (around 45 mg/l). At 24 h, concentrations at the depth of injection had reduced to around 13 and 15 mg/l for the drip and shank applications, respectively. At 48 h, concentration profiles were nearly uniform (at around 3 and 7 mg/l for the drip and shank applications, respectively), and at 168 h concentrations in both treatments were very low (0.7 mg/l). Generally, the MeI moved rapidly through the soil columns, particularly through the upper 20 cm of soil where the bulk density was lower (1.29 g/cm3) than in the deeper soil (1.61 g/cm3). Nevertheless, MeI was detected at maximum column depth (70 cm) in both treatments at 2 h after fumigation. Movement of the fumigant was most marked in the shank-injected treatment. The authors attributed this to the greater presence of water in the dripapplied treatment, which would have both maintained the MeI in the aqueous phase and physically limited its gaseous diffusion through the soil pore space. Moreover, they suggest that shank injection would therefore result in more effective gas-phase pest control. Also using a column approach, Gan et al. (1997) studied the soil distribution of shank-injected (30 cm depth) MeI and MeBr in relation to soil type and plastic tarp treatments. Comparing the two fumigants, these authors found that MeI behaved to a great extent like MeBr. However, they also noted that under the same conditions MeBr spread more rapidly throughout the soil column than MeI, probably due to its greater diffusion coefficient. Tarping of the soil clearly increased fumigant concentrations near the

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Figure 108.7  Time-wise trend in MeI gas distribution throughout 60-cm bare and HDPE tarped soil columns.

soil surface, when compared to a control (nontarped) soil. The authors suggest that this is likely to increase the fumigant effectiveness under tarped conditions. It is interesting that despite maintaining MeI within the soil, the polyethy­ lene tarp these authors used did not reduce overall MeI emissions by more than a few per cent, suggesting that it provided only a short-term barrier to MeI transfer from soil to air. Gan et al. (1997) attributed differences in the distribution of MeI between soil types to the differing soil water contents, bulk densities, degradation rates, and adsorption coefficients of the soils. They conclude that from the perspective of efficacy, the same dosage of MeI (or MeBr) in an organic matter-rich soil might not produce the same control as in an organic matter-poor soil due to increased fumigant degradation at higher organic matter contents. Despite an absence of investigations of MeI emissions under field conditions, the distribution of MeI within field soil has been determined. Using 3  3-m field plots, Gan et al. (1997) studied the time-dependent soil profile distribution of MeI and MeBr gas applied at a 30-cm depth in a sandy loam soil under polyethylene tarp. In concurrence with their work using soil columns (described above), these authors found that, initially, diffusion of MeI was more limited than MeBr. In the period of a few hours to a few days after injection, the concentrations of both fumigants decreased rapidly with time, becoming very small

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after 72 h. Although MeI continued to exhibit a more limited, slower diffusion over this period, its concentrations increased to become greater than those of MeBr around the depth of fumigant placement. Additionally, over this period the plastic tarp appeared to be ineffective in maintaining elevated fumigant concentrations in the surface soil, probably as a result of increased tarp permeability at higher temperatures occurring outdoors. Beyond 120 h, concentrations of MeI became higher than MeBr at most depths. This was attributed to the longer half-life of MeI in soil (Gan and Yates, 1996) and suggests that MeI may provide improved pest control compared to MeBr. Gan et al. (1997) suggested that the greater persistence of MeI in soil may increase the risk of leaching to groundwater.

108.6.2.5  Leaching of MeI Residues In reviewing the behavior of MeI, Ruzo (2006) concluded that MeI use was not a potential groundwater problem. However, Guo et al. (2004) suggested that although the proportion of MeI leached from experimental soil columns was small, the risk associated with leaching should be considered in view of groundwater protection. These workers studied the leaching of MeI for 19 days after drip or shank application, during which time emissions of the fumigant from the surface of the soil columns were prevented by sealing the surface with a virtually impermeable film. Of the approximately 228 mg of MeI added to their systems, 37.5 and 21.4 g were leached out of the shank and drip application treatments, respectively (0.02%). Concentrations of MeI in the column leachate were initially high (15.0 and 9.9 g/l for shank and drip applied treatments, respectively) and decreased rapidly in the first pore volume. This was followed by a long tailing of the curve up to four pore volumes. Up to around three pore volumes, MeI concentrations of the leachate were greater in the shank-injected treatment. Guo et al. (2004) attributed this to the likely higher soil moisture content in the drip-applied treatment, which may have reduced the sorption of the fumigant onto the soil solids. Alternatively, they suggest, the higher gas phase concentrations observed in the shank-injected treatment may have led to a greater amount of MeI becoming entrapped in soil intra-aggregate pores, where it is resistant to degradation and volatilization but susceptible to leaching (Guo et al., 2003). Clearly, there exists a small potential for MeI leaching with the liquid phase of soil.

Conclusion As a preplant soil fumigant, methyl iodide appears to be an effective potential alternative to methyl bromide. While offering the environmental advantage of not being reactive with stratospheric ozone, existing research suggests that MeI is comparable to, and often more efficacious than, MeBr to a wide range of weed, nematode, and microbial

pathogens. However, its time-limited registration (and nonregistration in certain states, e.g., California) reflect its potential for environmental impact. Its propensity to convert to the gaseous phase (high Henry’s constant) and relatively high vapor pressure ensure that MeI diffuses rapidly through soil and is readily emitted from soil to air, even in the presence of surface containment such as plastic tarps. Coupled with its relatively long half-life, the potential exists for high emissions over more extended periods than might be expected for fumigants with shorter half-lives. It is apparent that there is a need for field studies to more realistically determine the flux dynamics of MeI emissions from soil over time. Nevertheless, with our current understanding, there is some concern over the use of MeI as an agricultural fumigant, particularly in relation to its exposure to agricultural workers and local populations during, and in the days following, a fumigation event. Although reducing this exposure is implicit in the guidelines covering the use of MeI (e.g., limits on application rates, minimum distances from occupied structures), a number of serious potential human health risks associated with its use have been identified and form the basis of opposition to the widespread use of MeI within the United States.

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Chapter | 108  Methyl Iodide

Guo, M., Zheng, W., Papiernik, S. K., and Yates, S. R. (2004). Distribution and leaching of methyl iodide in soil following emulated shank and drip application. J. Environ. Qual. 33, 2149–2156. Hemwall, J. B. (1959). A mathematical theory of soil fumigation. Soil Sci. 88, 184–190. Hutchinson, C. M., McGiffen, M. E., Ohr, H. D., Sims, J. J., and Becker, J. O. (1999a). Efficacy of methyl iodide soil fumigation for control of Meloidogyne incognita, Tylenchulus semipenetrans and Heterodera schachtii. Nematology 1, 407–414. Hutchinson, C. M., McGiffen, M. E., Ohr, H. D., Sims, J. J., and Becker, J. O. (1999b). Evaluation of methyl iodide as a soil fumigant for rootknot nematode control in carrot production. Plant Dis. 83, 33–36. Hutchinson, C. M., McGiffen, M. E., Ohr, H. D., Sims, J. J., and Becker, J. O. (2000). Efficacy of methyl iodide and synergy with chloropicrin for control of fungi. Pest Manag. Sci. 56, 413–418. Hutchinson, C. M., McGiffen, M. E., Sims, M. E., Sims, J. J., and Becker, J. O. (2004). Fumigant combinations for Cyperus esculentum L control. Pest Manag. Sci. 60, 369–374. Ibekwe, A. M., Grieve, C. M., and Yang, C. H. (2007). Survival of Escherichia coli 0157:H7 in soil and on lettuce after soil fumigation. Can. J. Microbiol. 53, 623–635. Kabir, Z., Fennimore, S. A., Duniway, J. M., Martin, F. N., Browne, G. T., Winterbottom, C. Q., Ajwa, H. A., Westerdahl, B. B., Goodhue, R. E., and Haar, M. J. (2005). Alternatives to methyl bromide for strawberry runner plant production. Hortscience 40, 1709–1715. Ohr, H. D., Sims, J. J., Grech, N. M., Becker, J. O., and McGiffen, M. E. (1996). Methyl iodide, an ozone-safe alternative to methyl bromide as a soil fumigant. Plant Dis. 80, 731–735. Parkin, T. B., and Shelton, D. R. (1994). Modeling environmental effects of enhanced carbofuran degradation. Pestic. Sci. 40, 163–168. Ruzo, L. O. (2006). Physical, chemical and environmental properties of selected chemical alternative for the pre-plant use of methyl bromide as a soil fumigant. Pest Manag. Sci. 62, 99–113. Schneider, S. M., Ajwa, H., and Trout, T. J. (2006). Chemical alternatives to methyl bromide for nematode control under vineyard replant conditions. Am. J. Enol. Viticult. 57, 183–193. Sims, J. J., Grech, N. M., Becker, J. O., McGiffen, M. and Ohr, H. D. (1995). Methyl iodide: a potential alternative to methyl bromide. In “Proceedings of the Second Annual International Research Conference on Methyl Bromide Alternatives and Emissions Reductions, San Diego, 6-8 November 1995,” p. 46. Methyl Bromide Alternatives Outreach: Fresno, CA.

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Stanghellini, M. E., Ferrin, D. M., Kim, D. H., Waugh, M. M., Radewald, K. C., Sims, J. J., Ohr, H. D., Mayberry, K. S., Turini, T., and McCaslin, M. A. (2003). Application of preplant fumigants via drip irrigation systems for the management of root rot of melons caused by Monosporascus cannonballus. Plant Dis. 87, 1176–1178. Stromberger, M. E., Klose, S., Ajwa, H., Trout, T., and Fennimore, S. (2005). Microbial populations and enzyme activities in soils fumigated with methyl bromide alternatives. Soil Sci. Am. J. 69, 1987–1999. Unruh, J. B., Brecke, B. J., Dusky, J. A., and Godbehere, J. S. (2002). Fumigant alternatives for methyl bromide prior to turfgrass establishment. Weed Technol. 16, 379–387. U.S. EPA. (2008). “Time-limited Registration of Iodomethane (Methyl Iodide)”. Available at: http://www.epa.gov/pesticides/factsheets/ iodomethane_fs.htm (Oct 9 2008), Accessed Jan 6 2009. Walker, A., Moon, Y. H., and Welch, S. J. (1992). Influence of temperature, soil moisture and soil characteristics on the persistence of alachor. Pestic. Sci. 35, 109–116. Wang, D., and Yates, S. R. (1999). Spatial and temporal distributions of 1,3-dichloropropene in soil after drip and shank application and implications for pest control efficacy using concentration-time index. Pestic. Sci. 55, 154–160. Webster, T. M., Csinos, A. S., Johnson, A. W., Dowler, C. C., Sumner, D. R., and Fery, R. L. (2001). Methyl bromide alternatives in a bell pepper-squash rotation. Crop Prot. 20, 605–614. Yates, S. R., and Gan, J. (1998). Volatility, adsorption and degradation of propargyl bromide as a soil fumigant. J. Agric. Food Chem. 46, 755–761. Zhang, W. M., McGiffen, M. E., Becker, J. O., Ohr, H. D., Sims, J. J., and Campbell, S. D. (1998). Effect of soil physical factors on methyl iodide and methyl bromide. Pestic. Sci. 53, 71–79. Zhang, W. M., McGiffen, M. E., Becker, J. O., Ohr, H. D., Sims, J. J., and Kallenbach, R. L. (1997). Dose response of weeds to methyl iodide and methyl bromide. Weed Res. 37, 181–189. Zheng, W., McGiffen, M. E., Becker, J. O., Ohr, H. D., Sims, J. J., and Campbell, S. D. (1998). Effect of soil factors on methyl iodide and methyl bromide. Pestic. Sci. 53, 71–79. Zheng, W., Papiernik, S. K., Guo, M., and Yates, S. R. (2003). Accelerated degradation of methyl iodide by agrochemicals. J. Agric. Food Chem. 51, 673–679. Zheng, W., Papiernik, S. K., Guo, M., and Yates, S. R. (2004). Remediation of methyl iodide in aqueous solution and soils amended with thiourea. Environ. Sci. Technol. 38, 1188–1194.

Index

A

Abamectin, see also Avermectins skin testing, 749 Absorption, see also Percutaneous absorption; specific pesticides definition, 4 human study measurements carbaryl, 1284 carcinogens, 1284 DDT, 1283–1284 difficulties, 1283 herbicides, 1284 overview, 1282–1283 parathion, 1284 overview, 877, 880–881 membrane bilayer, 878 transport mechanisms passive transport, 878–879 specialized transport, 879 gastrointestinal absorption, 886–887 respirator y tract absorption, 887–888 injection routes, 888 prospects for study, 888–889 species differences dermal absorption, 84–85 gastrointestinal absorption, 83–84 ACD, see Allergic contact dermatitis Acephate, skin testing, 735 Acetochlor properties, 1757–1758 skin testing, 768 structure, 1754 toxicity in animals, 1758–1760 uses, 1758 Acetofenate properties, 2016 structure, 1976 toxicity, 2016 uses, 2016 Acetylcholine receptor imidacloprid interactions, 809–811 pyrethroid interactions, 806 Acetylcholinesterase age differences in organophosphorous pesticide toxicity, 824–825 antidotes, 1472–1473 assay colorimetric assay, 1466–1468 electrometric assay, 1465–1466 field kits, 1469–1470 radioassay, 1465 standards, 1469 variability, 1468–1469 biomarker for risk assessment, 1473 blood levels and exposure detection, 1471–1472

catalytic mechanism, 1461–1462 distribution, 1457–1458 functional overview, 1545 inhibitor toxicity, 1463–1464 reactivation of inhibited enzyme, 1471 structure, 1459–1461 substrate specificity and inhibitors, 1459 Acifluorfen, see Protoporphyrinogen oxidase-inhibiting herbicides Action, chemical, 9 Activated charcoal, poisoning management, 1297–1298 Acute exposure, risk assessment examples methyl bromide, 344–349 methyl parathion, 340–344 exposure databases, 339–340 Federal Insecticide, Fungicide, and Rodenticide Act guidelines, 349–350 toxicological databases, 337–338 Acute poisoning, see Poisoning Adicarb, skin testing, 739–740 Adjuvants, skin testing, 781 Aesthetic injury level (AIL), pests, 258 Age differences, pesticide toxicity animal toxicity testing, 71–72, 849–850 carbamates, 828 child health and regulation of pesticides, 821–822 general factors, 819–821 miscellaneous pesticides, 828–830 organochlorine insecticides, 826–827 organophosphorous pesticides, 822–826 pyrethroids, 828, 1670 Agricultural Handlers Exposure Database (AHED), 1149 Agricultural Health Study (AHS), 1364 Agricultural Reentry Task Force (ARTF), transfer coefficient database, 1150–1153 AHED, see Agricultural Handlers Exposure Database AHS, see Agricultural Health Study AIL, see Aesthetic injury level Alachlor carcinogenicity mechanisms, 1765–1768 percutaneous absorption, 882–883 properties, 1753 skin testing, 768–769 structure, 1754 toxicity animals, 1753–1757 humans, 1757 uses, 1753

Aldicarb acute toxicity, 1620 age differences in toxicity, 828 applications, 1619 carcinogenicity, 1622 cholinesterase inhibition pharmacokinetics, 1623–1624 chronic toxicity, 1622 critical effect and point of departure, 1624 developmental and reproductive toxicity, 1621 exposure food consumption data, 1624–1625 scenario modeling, 1626–1627 water consumption data, 1625–1626 genotoxicity, 1621–1622 human volunteer studies, 1622–1623 immunotoxicity, 485 infant and child sensitivity, 1621 mechanism of action, 1619 neurotoxicity, 1620–1621 risk characterization food, 1627 food and water, 1628–1629 prospects, 1629–1630 water, 1628 Aldrin developmental and reproductive toxicity, 407 overview, 2066–2071 Allergic contact dermatitis (ACD) agrochemical induction prevalence, 670 chemicals and cases, 670–672 diagnosis and treatment, 672–673 epidemiology, 669 pathophysiology, 669–670 prevention, 673–674 Allergy, dosage–response relationships, 50 Allethrin metabolism, 1636 properties, 1636 skin testing, 743 Allidochlor, skin testing, 769 Altitude, toxicity testing influences, 79 Aluminum phosphide, skin testing, 764 Ametryn, see also Symmetrical triazine herbicides Amidases, biotransformation, 870 Amitraz developmental and reproductive toxicity, 407 skin testing, 750 veterinary toxicology, 295

2319

Index

2320

Amitrole, developmental and reproductive toxicity, 397 Anabasine animal toxicity, 134 historical perspective, 134 human poisoning and treatment, 134–135 mechanism of action, 134 physicochemical properties, 134 stability, 134 Analytical chemistry, see Residue analysis Androgen receptor, pesticide effects on function linuron, 554 procymidone, 555 prodione, 555 vinclozolin, 554–555 Anilazine, skin testing, 757 Animal models age differences in sensitivity, 849–850 dose, timing and duration, 849 endpoints electrophysiology, 850 imaging, 850–851 omics approaches, 850 nonhuman primates Cambridge Neuropsychological Automated Test Battery, 854–855 electroencephalography, 851–852 National Center for Toxicological Research Operant Test Battery, 852–855 neonatal and infant assessments, 851 relevance to humans, 847–848 types, 848 Antimicrobial agents, skin testing, 706–710, 734–735 ANTU, see -Naphthylthiourea Area under the curve (AUC), pharmacokinetics, 926 ARTF, see Agricultural Reentry Task Force Arthropods, see Vector-borne disease Asthma cockroach induction, 259 ozone exacerbation, 579–580 Atrazine, see also Symmetrical triazine herbicides hypothalamic–pituitary–gonadal axis disruption, 558–559 immunotoxicity, 491–492 skin testing, 775 Atropine, pesticide poisoning management, 1302, 1557–1558 AUC, see Area under the curve Avermectins chemistry and formulations, 2093–2095 hazard identification and dose response, 2096–2101 ivermectin exposure in humans, 2101–2102 mechanism of action, 2096 overview, 2093 placental P-glycoprotein in protection, 2106–2107 risk characterization, 2102–2104 uses, 2095–2096 Avitrol, bird flock dispersal, 283 Azadirachtin

animal toxicity acute, 151–153 cytotoxicity, 153–154 irritation, 153 formulations and uses, 149–150 human poisoning and treatment, 154 mechanism of action, 150–151 metabolism and excretion, 151 neem tree uses, 147–148 physicochemical properties, 149 reproduction effects, 151 skin testing, 749 stability, 149 structure, 149 Azafenidin, see Protoporphyrinogen oxidase-inhibiting herbicides

B

Bacillus cereus, microbial pest control and toxicity, 451 Bacillus sphaericus, microbial pest control and toxicity, 451–453 Bacillus thuringiensis (Bt) animal toxicity acute, 158–160 chronic, 158 -exotoxin, 161 genotoxicity, mutagenicity, and cytotoxicity, 160 irritation, 158 poisoning, 160–161 sensitization, 158 distribution and excretion, 157–158 formulations and uses, 155–156 history of use, 154 human toxicity irritation and sensitization, 161 poisoning and treatment, 161–162 mechanism of action, 115, 156–157 microbial pest control and toxicity, 447–451 nomenclature, 154–155 skin testing, 748–749 stability, 156 structure of toxins, 155 Bacterial reverse mutation assay, genotoxicity testing, 360–361 Baculovirus, microbial pest control and toxicity, 456 Baits bird poison bait, 283–284 insect control, 265–266 rodent baits, 277–281 BBA model, occupational exposure assessment, 1146–1149 Behavior, see also Neurotoxicity assessment of behavioral changes, 587–589 organophosphate toxicity human effects, 837–839 mouse models, 840–842 rat models, 839–840 zebrafish models, 843 Bendiocarb developmental and reproductive toxicity, 412 skin testing, 740 Benefin, skin testing, 770

Benomyl developmental and reproductive toxicity, 415 skin testing, 754–755 Bensulide, skin testing, 771 Bentazon, genotoxicity testing, 367 1,2-Benzisothiazol-3(2H)-one, skin testing, 706–707 Bergamote, coumarins as rodenticides, 2205–2206 Bifenox, see Protoporphyrinogen oxidase-inhibiting herbicides Bifenthrin metabolism, 1637–163 properties, 1636–1637 skin testing, 743–744 Bilanafos animal toxicity, 183 formulations and uses, 183 human toxicity, 183–184 mechanism of action, 183 metabolism and excretion, 183 physicochemical properties, 183 stability, 183 structure, 183 Bioallethrin, age differences in toxicity, 828 Bioavailability, pharmacokinetics, 927 Biomonitoring applications exposure and dose estimation forward dosimetry, 1026 reverse dosimetry, 1026–1027 population-based exposure trend assessment, 1022–1026 biomarkers biological relevance, 1029 examples, 1028–1029 excretory pesticide exposure biomarkers, 965–966 kinetics, 1032–1031 practicality, 1029–1032 sensitivity, 1027 specificity, 1027 urinary excretion rate, 1031–1033 validity, 1029 genotoxicity testing, 369–373 percutaneous absorption, 696 pesticide exposure linking to health effects, 1021–1022 prospects for study, 1033–1036 sample collection for exposure assessment, 986 turf pesticide exposure, 1042, 1067 Biotransformation amidases, 870 antagonism, 950–951 biphasic inhibition and induction by pesticides, 951–952 central nervous system, 909 chlorpyrifos, 1507–1510 cytochromes P450, 866–867, 894–899, 900–906 DDT, 1982–1986 DDT-dechlorochlorinase, 870–871 enzyme assays, 897–898 enzyme–pesticide interactions

Index

activation of enzymes, 953 induction of enzymes mechanisms, 948 microsomal enzymes, 941–948 miscellaneous enzymes, 948 inhibition of enzymes distribution and blood levels, 949 inhibition of metabolism of other pesticides, 951 in vitro effects, 949–950 in vivo effects, 948–949 epoxide hydrolase, 869, 899 ethylenebisdithiocarbamates, 1690–1691 external transformations, 893–894 factors affecting development, 910 genetics, 911 individual differences, 910 sex, 910–911 species, 910 flavin-containing monooxygenase, 867–869, 894, 898, 906 gastrointestinal tract, 909 hydrolases, 870 kidney, 908–909 lung, 908 malathion, 1529–1530 metabolism definition, 893 metabolite toxicity, 55–56, 909–910 nasal tissues, 908 organophosphorous insecticides conjugations, 1404 hydrolysis, 1402–1404 oxidations, 1400–1402 reductions, 1402 overview, 865–866 paraquat, 1779–1780, 1806 phase II conjugation reactions, 871–873, 899 potentiation, 950 prostaglandin synthetase, 869–870 protoporphyrinogen oxidase-inhibiting herbicides, 1746–1747 resistance, 911–912 skin, 908 species differences comparisons, 899, 907 phase I, 85–86 phase II, 86–87 synergism, 950 tolerance, 911 Bird control flock dispersal agent, 283 poison bait, 283–284 Blasticidin-S animal toxicity, 175 formulations and uses, 174 human toxicity, 175–176 mechanism of action, 174–175 metabolism and excretion, 175 physicochemical properties, 173–174 stability, 174 structure, 174 Blom equation, 1208 Blood–brain barrier, pesticide permeability, 924–925

2321

BLS, see Bureau of Labor Statistics Borates animal toxicity acute toxicity, 2040 chronic toxicity, 2040–2041 biological importance animals, 2037 humans, 2037–2038 plants, 2036–2037 boric acid, 750, 2035 boron equivalents, 2034 exposure dietary, 2035 environmental, 2036 occupational, 2035 residential, 2035–2036 human toxicity epidemiology, 2044–2048 occupational exposure, 2043–2044 overview, 2033–2034 poisoning animals, 2042 humans, 2042–2043 sodium salts, 2035 toxicokinetics absorption, 2038 distribution, 2038–2039 excretion, 2039–2040 metabolism, 2039 uses, 2034–2035 zinc borate, 2035 Brodifacoum animal toxicity, 2185–2187 human toxicity, 2187–2188 overview, 2171–2173 poisoning treatment, 2187 properties, 2184 structure, 2172 uses, 2184–2185 Bromacil, skin testing, 776 Bromadiolone animal toxicity, 2192–2193 human toxicity, 2193 overview, 2171–2173 poisoning treatment, 2193 properties, 2192 structure, 2172 uses, 2192 Bromethalin animal toxicity, 2204–2205 human toxicity, 2205 poisoning treatment, 2205 properties, 2204 structure, 2200 uses, 2204 2-Bromo-2-nitropropane-1,3-diol, skin testing, 734–735 Bromoxynil developmental and reproductive toxicity, 397 skin testing, 777–778 Bt, see Bacillus thuringiensis Buehler test, skin testing, 702 Bureau of Labor Statistics (BLS) case definition, 1339 data sources, 1339

description, 1339 findings, 1341–1342 information sources, 1364 period of data collection, 1339 report frequency, 1339 target population, 1339 Burkholderia cepacia, microbial pest control and toxicity, 453–454 Butachlor carcinogenicity mechanisms, 1766–1768 properties, 1760 skin testing, 769 structure, 1754 toxicity in animals, 1760–1762 uses, 1760 Butafenacil, see Protoporphyrinogen oxidase-inhibiting herbicides Butoxypolypropylene glycol, skin testing, 750 Butyrylcholinesterase assay, 1467–1468 blood levels and exposure detection, 1471 catalytic mechanism, 1461–1462 deficiency, 1465 distribution, 1457–1458 gene polymorphisms and pesticide susceptibility, 539 structure, 1459–1461 substrate specificity and inhibitors, 1459

C

C/A index, cumulative effect measurements, 24–26 Calcineurin, pyrethroid neurotoxicity modulation, 806 Calcium channel, modulation by pyrethroids, 805, 1677–1679 Calendex, dietary exposure assessment, 1112–111 California Environmental Protection Agency, 1364 Cambridge Neuropsychological Automated Test Battery, nonhuman primate studies, 854–855 Cancer see Carcinogenesis; Carcinogenicity testing Capillary electrophoresis (CE), pesticide residue analysis, 1123–1124 Captafol mechanism of toxicity, 1938 skin testing, 753–754 structure, 1938 Captan history and use, 1016 human risk assessment, 1939–1940 metabolism, 1920–1922 overview, 1915–1916 physical properties and chemical reactions, 1917–1920 skin testing, 753 toxicology acute toxicity, 1922–1924 carcinogenicity, 1933–1937 chronic toxicity, 1926–1927 developmental toxicity, 1927–1928 mechanism of toxicity, 1937–1938

Index

2322

Captan (Continued) mutagenicity, 1928–1933 overview, 1916–1917 reproductive toxicity, 1928 subchronic toxicity, 1924–1926 Carbaryl absorption measurements in humans, 1284 acute toxicity, 1607–1608 age differences in toxicity, 828 applications, 1607 carcinogenicity mouse, 1610–1612 rat, 1610 developmental and reproductive toxicity, 413, 1609–1610 genotoxicity, 1610 mechanism of action, 1607 metabolism in rat, 1612–1614 neurotoxicity, 1608–1609 skin testing, 740 subchronic toxicity, 1608 toxicokinetics in rat, 1614–1615 Carbofuran developmental and reproductive toxicity, 412 physiologically-based pharmacokinetic modeling, 1598–1599 Carbon disulfide, see Soil fumigants Carboxin, skin testing, 757–758 Carboxyl esterases, gene polymorphisms and pesticide susceptibility, 538–539 Carcinogenesis chronic poisoning risks, 1308–1309 gene–pesticide interactions, 526 Carcinogenicity testing, see also specific pesticides challenges cumulative toxicity, 480 life stages, 479–480 dosage–response relationships, 48–50 Environmental Protection Agency carcinogenicity classification, 476 human relevance, 477–478 overview, 474–475 pesticide carcinogenesis mechanisms, 477 risk assessment noncancer endpoints, 478–479 nonthreshold carcinogens, 479 threshold carcinogens, 479 statistical analysis, 475, 477 CARES, see Cumulative and Aggregate Residue Evaluation System Carfentrazone, see Protoporphyrinogen oxidase-inhibiting herbicides Case-control study, 1361–1362 Cathartics, poisoning management, 1298–1299 Cayanazine, skin testing, 775 CCA, see Chromated copper arsenate CDAS, see Customized Dietary Assessment Software CDF, see Cumulative distribution function CE, see Capillary electrophoresis Cevadine animal toxicity acute, 144 teratogenicity, 144

formulations and uses, 142–143 human toxicity irritation, 145 poisoning, 145 treatment, 145 mechanism of action, 143 metabolism and excretion, 144 pharmacology, 143–144 structure, 142 Chloracetanilide herbicides, see Acetochlor; Alachlor; Butachlor; Metolachlor; Propachlor Chloralose animal toxicity, 2200–2201 human toxicity dosage response, 2202 incidents, 2201–2202 laboratory findings, 2202 therapeutic use, 2201 poisoning treatment animals, 2201 humans, 2202 properties, 2200 structure, 2200 uses, 2200 Chlorantraniliprole acute toxicity, 2233, 2235 adrenal function effects, 2237–2238 chronic toxicity, 2237 developmental toxicity, 2238–2239 genotoxicity, 2240 human toxicity, 2240–2241 immunotoxicity, 2240 mechanism of action, 2232 metabolism, 2233–2234 neurotoxicity, 2239 overview, 2231 properties, 2231–2232 regulatory reference values, 2241–2242 structure, 2231 subchronic toxicity, 2235–2237 toxicokinetics dermal administration, 2233 oral administration, 2232–2233 uses, 2232 Chlordane, overview, 2071–2073 Chlordecone age differences in toxicity, 827 developmental and reproductive toxicity, 409 overview, 2075–2076 Chlordene, overview, 2071–2073 Chlordoimeform age differences in toxicity, 829 Chloride channel, modulation by pyrethroids, 808, 1679–1680 Chlorobenzilate animal toxicity, 2015 human toxicity, 2015 properties, 2014 structure, 1976 uses, 2015 5-Chloro-2-methyl-4-isothiazoline, skin testing, 706–707 Chloroneb, skin testing, 758

Chlorophacinone animal toxicity, 2188–2189 human toxicity, 2189–2190 overview, 2171–2173 poisoning treatment, 2189–2190 properties, 2188 structure, 2172 toxicokinetics, 2189 uses, 2188 Chloropicrin, see also Soil fumigants adjustment factors, 575 skin testing, 766 Chlorothalonil acute toxicity, 1953–1955 carcinogenicity, 1958–1960 chronic toxicity, 1957–1958 developmental toxicity, 1960–1961 genotoxicity, 1958 human toxicity eye, 1962 poisoning, 1963 respiratory effects, 1962–1963 skin, 1962 investigative toxicity studies, 1961 properties, 1951 reproductive toxicity, 1961 sensitization, 1955 structure, 1951 subchronic toxicity, 1955–1957 toxicokinetics dermal administration, 1952 oral administration, 1951–1952 uses, 1951 Chlorothalonil, skin testing, 758–759 Chlorpyrifos absorption, 1506–1507 acute and subchronic toxicity, 1512–1514 age differences in toxicity, 825 behavior effects in development, 839–843 biotransformation, 1507–1510 carcinogenicity, 1515–1516 chronic toxicity, 1514–1515 clothes residuals after washing, 693 developmental and reproductive toxicity, 410–411, 1516–1518 distribution, 1507 excretion, 1510 exposure, 1510–1511 exposure assessment, 1013–1016 genotoxicity, 367, 1515–1516 immunotoxicity, 489 mechanism of action, 1511–1512 physicochemical properties, 1505 regulation, 1505 skin testing, 735–736 toxic interactions, 1518–1519 turf pesticide exposure, 1038, 1040 Chlorsulfuron, skin testing, 776 Cholecalciferol animal toxicity, 2198–2199 human toxicity, 2199 properties, 2198 toxicokinetics, 2198–2199 uses, 2198

Index

Cholinesterases, see also Acetylcholinesterase; Butyrylcholinesterase comparative sensitivity studies in neurotoxicity, 599 inhibitor classification, 292 Chromated copper arsenate (CCA), exposure assessment in children, 1013 Chronic poisoning, see Poisoning Chronic toxicity testing administration route, 466 chemical purity, 467–468 dose levels, 466–467 endpoints body weight, 469 clinical chemistry, 471–472 clinical observations, 468–469 eye, 469 hematology, 470–471 mortality, 468 organ weight, 472–473 overall assessment, 473–474 overview, 468 pathology, 473 urinalysis, 472 purpose, 463 regulatory requirements, 464–466 species and study duration, 464, 466 Chronicity index, cumulative effect measurements, 26–27 Cinerin, see Pyrethrins Circadian rhythm, chemical toxicity testing influences, 81–82 Cl, see Plasma clearance CLOGP program, hydrophobicity estimation, 1237–1240 Cloransulam-methyl properties, 1866 structure, 1865 toxicity animals, 1866–1868 humans, 1868 uses, 1866 Clotting, chronic toxicity testing, 471 Cluster analysis, percutaneous absorption modeling, 692, 694 Codlemone, see Pheromones Cohort study, 1361 Communication, see Risk communication Conditioned position responding, National Center for Toxicological Research Operant Test Battery, 852 Contact urticaria syndrome (CUS) definition, 677 diagnosis, 679 etiology and mechanisms immunologic contact urticaria, 678 nonimmunologic contact urticaria, 678 uncertain mechanism, 678–679 pesticide induction, 679–680 predictive assays, 679 signs and symptoms, 677–678 staging, 677 Contraceptives, vertebrate pest control, 282 Copper (II) hydroxide, skin testing, 756–757

2323

Copper naphthenate, skin testing, 757 Coumafuryl, overview, 2182 Coumaphos, skin testing, 736 Cross-sectional study, 1362 Crowding, toxicity testing influences, 74–75 Cumulative and Aggregate Residue Evaluation System (CARES), dietary exposure assessment, 1113 Cumulative distribution function (CDF), sensitivity and uncertainty analysis, 1012 Cumulative effect measurements C/A index, 24–26 chronicity index, 26–27 historical perspective, 23–24 Curcurbitacins, see Pheromones CUS, see Contact urticaria syndrome Customized Dietary Assessment Software (CDAS), dietary exposure assessment, 1114 Cyanazine, see also Symmetrical triazine herbicides Cyanuric acid, skin testing, 707 Cyclodienes reductive chlorination, 207 resistance mechanisms, 2083 Cycloprothrin metabolism, 1638 properties, 1638 Cyfluthrin metabolism, 1639 properties, 1638 skin testing, 745 Cyhalothrin metabolism, 1639–1640 properties, 1639 skin testing, 745 Cypermethrin age differences in toxicity, 828 metabolism, 1640–1641 properties, 1640 skin testing, 746 Cyphenothrin metabolism, 1641–1642 properties, 1641 Cyprodinil animal toxicity acute toxicity, 1906 chronic toxicity, 1907 lipoprotein response, 1908–1909 mutagenicity, 1911 neurotoxicity, 1910–1911 pharmacological studies, 1911 reproductive and developmental toxicity, 1909–1910 subchronic toxicity, 1906–1907 human toxicity, 1911–1912 mechanism of action, 1904 metabolism, 1904–1905, 1907–1908 overview, 1903, 1912 poisoning treatment, 1912 properties, 1904 structure, 1903 synonyms, 1903

toxicikinetics goat, 1905 hen, 1905–1906 rat, 1904 uses, 1904 CYPs, see Cytochromes P450 Cytochromes P450 (CYPs) activation by pesticides, 953 biotransformation, 866–867, 898–899 gene polymorphisms and pesticide susceptibility, 529–536 hepatotoxicity, 953 induction by pesticides mechanisms, 948 microsomal enzymes, 941–948 organophosphorous insecticide metabolism, 1400–1401 pesticide-specific reactions, 900–906 reaction types, 894–897

D

2,4-D, see 2,4-Dichlorophenoxyacetic acid Dazomet, adjustment factors, 575 2,4-DB, see 4-(2,4-Dichlorophenoxy) butyric acid DDT, see 1,1-(2,2,2-Trichloroethylidene)bis(4-chlorobenzene) DDT-dechlorochlorinase, biotransformation, 870–871 DEEM, see Dietary Exposure Evaluation Model DEET, see N,N-Diethyl-3-methylbenzamide Delayed-matching to sample task, National Center for Toxicological Research Operant Test Battery, 853 Deltamethrin age differences in toxicity, 828 metabolism, 1643–1644 properties, 1642–1643 skin testing, 746 Department of Pesticide Regulation (DPR), California case definition, 1333–1334–1335 data sources, 1334 description, 1332–1333 findings, 1334–1338 period of data collection, 1334 report frequency, 1334 target population, 1334 DEPM, see Dietary Exposure Potential Model DERM, see Dermal exposure ranking method Dermal assessment estimate method (DREAM), percutaneous absorption modeling, 695 Dermal exposure ranking method (DERM), percutaneous absorption modeling, 694–695 Dermatitis, see Allergic contact dermatitis; Irritant contact dermatitis; Percutaneous absorption; Photocontact dermatitis; Skin testing

Index

2324

Developmental toxicity, see also specific pesticides epidemiology, 383–384 exposure adult reproductive toxicants, 385 prenatal reproductive toxicants, 384–385 prepubertal reproductive toxicants, 385 timing, 384 fungicides benomyl, 415 dinocap, 415 ethyl dibromide, 419 folpet, 415 hexachlorobenzene, 415, 419 imazalil, 420 ketoconazole, 420 maneb, 420 metam sodium, 420–421 methyl thiophanate, 421 pentachlorophenol, 421 propioconazole, 420 table, 416–419 terrazole, 421 vinclozolin, 421–422 herbicides 2,4-D, 391, 397 amitrole, 397 bromoxynil, 397 dinoseb, 398 diquat, 398 diuron, 399 ethofumesate, 398 ethyl dipropyl thiocarbamate, 398 linuron, 399 milinate, 398 nitrofen, 399 paraquat, 398 table, 392–397 triazines, 399 insect growth regulators methoprene, 414 diflubenzuron, 414 fenoxycarb, 414–415 insecticides aldrin, 407 amitraz, 407 bendiocarb, 412 carbaryl, 413 carbofuran, 412 chlordecone, 409 chlorpyrifos, 410–411 DDT, 408 diazinon, 411 dibromochloropropane, 407–408 dimethoate, 411 fenthion, 411 lindane, 409 malathion, 411–412 methoxyxchlor, 408–409 overview, 399, 407 parathion, 411 pyrethrins, 413–414 table, 400–407 thiodicarb, 413 thiram, 413

trichlorfon, 412 zineb, 413 mechanisms of pesticide action, 385 methyl bromide, 427 overview, 382–383 rodenticides, 422 testing dose selection, 388 exposure assessment, 389–390 Food Quality Protection Act impact, 390 interpretation, 388–389 principles, 386–388 regulatory history, 386 species selection, 388 statistical evaluation, 389 veterinary pesticides, 422–427 Dialkylthiocarbamates, see Ethylenebisdithiocarbamates Diazepam, pesticide poisoning management, 1561 Diazinon age differences in toxicity, 822 behavior effects in development, 839–841 biotransformation, 1404 developmental and reproductive toxicity, 411 immunotoxicity, 488–489 skin testing, 736 Dibromochloropropane age differences in toxicity, 829 developmental and reproductive toxicity, 407–408 1,2-Dibromo-2,4-dicyanobutane, skin testing, 734–735 2,2-Dibromo-3-nitril-propionamide, skin testing, 734–735 Dicamba formulations, 1849 poisoning treatment, 1851 properties, 1949 structure, 1849 synonyms, 1849 toxicity animals, 1850 genotoxicity, 1850–1851 humans, 1850 reproductive toxicity, 1850 toxicokinetics, 1849–1850 uses, 1849 Dicamba, skin testing, 773 Dichlobenil, skin testing, 778 Dichlofluanid mechanism of toxicity, 1938–1939 structure, 1938 1,4-Dichlorobenzene, genotoxicity testing, 369 1,1-Dichloro-2,2-bis(4-chlorophenyl)ethane (TDE) animal toxicity, 2004–2007 human toxicity analytical findings, 2008 therapeutic use, 2007–2008 structure, 1976 Dichlorodiphenyltrichloroethane (DDT), see 1,1-(2,2,2-Trichloroethylidene) -bis(4-chlorobenzene) 2,4-Dichlorophenoxyacetic acid (2,4-D) absorption, 1832–1833

animal toxicity acute toxicity, 1834 carcinogenicity, 1839–1840 chronic toxicity, 1839–1840 genotoxicity, 1840 immunotoxicity, 1838 neurotoxicity, 1838–1839 reproductive and developmental toxicity, 1834–1838 subchronic toxicity, 1834–1835 developmental and reproductive toxicity, 391, 397 distribution, 1833 excretion, 1833–1834 forms, 1829 formulations, 1832 history of use, 1831 human exposure, 1832 human toxicity, 1840–1841 immunotoxicity, 490 metabolism, 1833 overview, 1829 pharmacokinetics, 1833 properties, 1829 skin testing, 772–773 structure, 1830 thyroid hormone disruption, 561 4-(2,4-Dichlorophenoxy) butyric acid (2,4-DB), properties, 1830 2-(2,4-Dichlorophenoxy) propionic acid (2,4-DP), properties, 1830 1,3-Dichloropropene, see also Soil fumigants acute toxicity, 2282 adjustment factors, 574–575 chronic toxicity, 2285–2286 dose response, 2286–2288 formulations, 2281 human toxicity accidental poisoning, 2288 experimental exposure, 2288 use experience, 2288–2289 metabolism, 2304 properties, 2281 repeated dose toxicity, 2282–2283 reproductive toxicity, 2283 risk characterization, 2289–2290 short-term assays, 2285 skin testing, 764–765, 2303–2304 structure, 2281 toxicokinetics, 2283–2285 uses, 2281 3,4-Dichloroproprionanilide, see Propanil Dichlorvos, skin testing, 736–737 Diclosulam properties, 1869 structure, 1865 toxicity animals, 1869–1870 humans, 1870 uses, 1869 Dicofiol animal toxicity, 2015 human toxicity, 2016 properties, 2015 structure, 1976 uses, 2015

Index

Dieldrin GABAA receptor interactions, 807–808, 2084–2086 overview, 2066–2071 reductive chlorination, 2078 Dienochlor, skin testing, 747 Dietary Exposure Evaluation Model (DEEM), dietary exposure assessment, 1113–1114 Dietary Exposure Potential Model (DEPM), dietary exposure assessment, 1114 Dietary exposure, see Exposure Dietary Record Generator (DRG), dietary exposure assessment, 1114 N,N-Diethyl-3-methylbenzamide (DEET) acute toxicity, 2112–2113 age differences in toxicity, 829 chemistry, 2111 chronic toxicity, 2115–2116 developmental toxicity, 2114–2115 genotoxicity, 2117 human toxicity neurological effects, 2121–2122 regulatory risk assessment, 2122–2123 skin reactions, 2121 neurotoxicity, 2116–2117 overview, 2111 percutaneous absorption enhancement of pesticides, 883 pharmacokinetics humans, 2118 interactions with chemicals environmental chemicals, 2120–2121 ethanol and sunscreens, 2119 rats, 2117–2118 reproductive toxicity, 2115 skin testing, 751 structure, 2112 subchronic toxicity, 2113–2114 vector-borne disease prevention, 247–248 Difenacoum animal toxicity, 2190–2191 human toxicity, 2191–2192 overview, 2171–2173 poisoning treatment animals, 2191 humans, 2192 properties, 2190 structure, 2172 uses, 2190 Difethialone animal toxicity, 2194 human toxicity, 2194 overview, 2171–2173 poisoning treatment, 2194 properties, 2194 structure, 2172 uses, 2194 Differential protection technique, exposure assessment, 1281–1282 Diflubenzuron, developmental and reproductive toxicity, 414 Dihydroheptachlor, overview, 2071–2073 Dimethoate developmental and reproductive toxicity, 411 skin testing, 737

2325

Dimethyl disulfide, see Soil fumigants Dinocap, developmental and reproductive toxicity, 415 Dinoseb, developmental and reproductive toxicity, 398 Diphacinone animal toxicity, 2183–2184 human toxicity, 2184 overview, 2171–2173 poisoning treatment, 2184 properties, 2183 structure, 2172 toxicokinetics, 2183–2184 uses, 2183 Diquat developmental and reproductive toxicity, 398 skin testing, 767–768 Disarture, see Pheromones Disease, toxicity testing influences, 76–77 Dissipation, see Environmental transport and fate Distribution blood–brain barrier, 924–925 overview, 923 placental transfer, 925 rate and extent, 924 species differences, 83, 85 storage and redistribution fat, 925 plasma proteins, 925 repeated exposures, 925–926 total body water, 924 volume of distribution, 924 Dithiobiuret, age differences in toxicity, 829 Dithiopyr, skin testing, 778 Diuron absorption, metabolism, and excretion, 1726 developmental and reproductive toxicity, 399 properties, 1726 skin testing, 776 structure, 1725 toxicity studies in animals, 1726–1727 uses, 1726 Dopamine transporter, gene polymorphisms and pesticide susceptibility, 540–541 Dose concentration times time concept, 4–8 definition, 6 testing considerations duration, 62–63 exposure route, 63–64 scheduling, 60–62 toxicity influences, 53–54 2,4-DP, see 2-(2,4-Dichlorophenoxy) propionic acid DPR, see Department of Pesticide Regulation DREAM, see Dermal assessment estimate method DRG, see Dietary Record Generator DTAM, skin decontamination, 698

E

EASE, see Estimation and assessment exposure EBDCs, see Ethylenebisdithiocarbamates EC 50, dispersed toxicants, 33

Ecologic study, 1362–1363 Economic injury level (EIL), pests, 258 Ecotoxicological risk assessment (ERA) conceptual models, 1199 ecosystem resiliency and redundancy, 1196–1197 effect characterization ecosystem, 1202–1203 individual species aquatic organisms, 1201 data sources, 1201–1202 organism groupings, 1202 terrestrial organisms, 1201 endpoints, 1196 exposure characterization estimation, 1204–1206 measurements, 1203–1204 extrapolations, 1197–1198 hazard quotient, 1207 measures of effect, 1196 multiple lines of evidence, 1212 overview, 1191–1192 pesticides in relation to other substances need for risk assessment, 1193 principles, 1193–1194 receptor organism characterization, 1194–1195 stressor characterization, 1194 system characterization, 1195 probabilistic risk assessment distributional approaches, 1211–1212 joint probability curve, 1209–1210 mixtures, 1210–1211 overview, 1207–1209 protection goals, 1195–1196 protoporphyrinogen oxidase-inhibiting herbicides, 1742–1745 risk communication, 1212–1213 scoring systems and criteria setting, 1206–1207 uncertainty, 1199 ED 01, study principles, 19–21 ED, 50 90-dose ED, 50 comparison with one-dose test, 18 determination, 19 one-dose ED, 50 confidence limits and reproducibility, 16 corresponding LD values, 15–16 curve shape, 13–15 determination, 16–17 dose volume, 17 overview, 12–13 EDB, see Ethylene dibromide EDCs, see Endocrine-disrupting chemicals EIL, see Economic injury level Electrophysiology animal study endpoint, 850 neurotoxicity assessment, 589–590 Elimination, see Excretion Ellman assay, acetylcholinesterase, 1466–1468 Emanectin benzoate, see Avermectins EMMs, see Exposure mitigation measures Empenethrin metabolism, 1644 properties, 1644

Index

2326

Endocrine-disrupting chemicals (EDCs), see also Reproductive toxicity androgen receptor function linuron, 554 procymidone, 555 prodione, 555 vinclozolin, 554–555 hypothalamic–pituitary–gonadal axis disruption atrazine, 558–559 dithiocarbamates, 558 esfenvalerate, 560 pyrethrins, 559–560 initiatives for impact analysis, 662 methoxychlor and estrogen receptor function, 553–554 steroid enhancers, 557 steroidogenesis inhibition ketoconazole, 555–556 molinate, 556–557 prochloraz, 557 testing multigenerational studies, 561–562 screening program, 562–565 thyroid hormone disruption 2,4-D, 561 ketoconazole, 560 triclosan, 561 Endosulfan acute toxicity, 501–502 age differences in toxicity, 827 biotransformation, 501–502 chronic toxicity, 503 developmental toxicity, 504 dietary exposure California tolerances, 512 DPR database, 510–511 highest measured acute residue values, 512–513 residue adjustments, 512 environmental fate, 500–501 Food Quality Protection Act safety factor, 515–516 genotoxicity testing, 367, 503 hazard identification acute toxicity, 505–506 chronic toxicity, 506 genotoxicity, 506 subchronic toxicity, 506 margins of exposure for risk characterization calculations, 514 outcomes, 514–515 overview, 513–514 mechanism of toxicity, 501 neurotoxicity acute, 504 developmental, 504 subchronic, 504–505 occupational exposure acute, 506–508 aerial application, 508–509 air aggregate exposure, 513 bystanders, 510 handlers, 508 reentry workers, 509–510

root dipping, 509 seasonal and annual, 507 overview, 2073–2075 pharmacokinetics, 501 physicochemical properties, 500 reproductive toxicity, 503–504 skin testing, 748 subchronic toxicity, 502 tolerance assessment, 516 Endothall, skin testing, 778 Endrin, overview, 2066–2071 Environmental Protection Agency (EPA) carcinogenicity classification, 476 exposure mitigation measures, 1159–1160 history of pesticide regulation, 1371–1372 information sources, 1365 pesticide risk assessment and human data utility, 1285–1290 pet exposure methodology activity patterns, 1087 formulation properties, 1086 overview, 1084–1086 postapplication dermal and incidental ingestion exposure variables, 1087 research needs, 1087–1089 risk information reporting requirements, 1350 Worker Protection Standard for Agricultural Pesticides, 1350–1351 Environmental transport and fate activation–deactivation, 1224 compartments, 1221–122 dissipation, 1221 modeling computer models, 1225 physical models, 1224–1225 overview, 1219–1221 pesticide structure influences, 1222–1224 Enzyme induction, dosage–response relationships, 50–51 EPA, see Environmental Protection Agency Epidemiologic studies case reports, 1363 case-control study, 1361–1362 cohort study, 1361 cross-sectional study, 1362 ecologic study, 1362–1363 principles of epidemiology, 1360 Epoxide hydrolase, biotransformation, 869, 899 EPTC, see Ethyl dipropylthiocarbamate ERA, see Ecotoxicological risk assessment Ergocalciferol animal toxicity, 2195–2196 human toxicity, 2197 poisoning treatment animals, 2196 humans, 2197 properties, 2195 uses, 2195 Ergodynamics, conceptualization, 8–9 Esfenvalerate hypothalamic–pituitary–gonadal axis disruption, 560 skin testing, 746 Estimation and assessment exposure (EASE), percutaneous absorption modeling, 695–696

Estrogen receptor, methoxychlor effects on function, 553–554 ET 50, principles, 27–28 Ethafluralin, skin testing, 770 Ethofumesate, developmental and reproductive toxicity, 398 Ethyl dibromide, developmental and reproductive toxicity, 419 Ethyl dipropylthiocarbamate (EPTC) developmental and reproductive toxicity, 398 exposure mitigation measures, 1167 Ethylan animal toxicity, 2008–2009 human toxicity, 2009–2010 properties, 2008 structure, 1976 uses, 2008 Ethylenebisdithiocarbamates (EBDCs) acute reference dose, 1702 acute toxicity, 1691–1692 carcinogenicity, 1702–1703 ethylthiourea metabolite equivalent doses, 1698–1699 no observable adverse effect level, 1700–1701 exposure assessment endpoints, 1702 genotoxicity, 1698 hazard characterization, 1700 hazard identification, 1689 human toxicology, 1703 liver effects, 1698 metabolism, 1690–1691 neurotoxicity, 1700 no observable adverse effect level, 1701–1703 reproductive and developmental toxicity, 1699–1700 risk characterization dietary intake, 1703–1704 occupational exposure, 1704 thyroid effects, 1691, 1695, 1697–1698 types, 1689 uses, 1689 Ethylene dibromide (EDB) genotoxicity testing, 365 skin testing, 765 Ethylene oxide, genotoxicity testing, 362, 365 Ethylthiourea, see Ethylenebisdithiocarbamates Etofenprox metabolism, 1644–1645 properties, 1644 Excretion see also specific pesticides alimentary elimination, 964 biliary excretion, 963 cellular elimination, 964–965 eggs, 964 fetus, 964 kidney function glomerular filtration, 962 tubular reabsorption, 962 tubular secretion, 962–963 milk, 964 obscure routes, 964 overview, 961 pesticide exposure biomarkers, 965–966 respiratory excretion, 963

Index

species differences urinary versus biliary, 86–87 urinary versus fecal, 87 Exposure, see also Occupational exposure assessment overview, 973–974 pesticide considerations, 974–976 biomonitoring, see Biomonitoring definition, 3–4, 971 dietary exposure aggregate exposure assessment, 1107, 1110–1112 calculations degradation calculation, 1109–1110 exposure methodology sequence, 1108–1109 Monte Carlo simulation, 1109 risk calculation, 1110 cumulative exposure assessment, 1106–1107 food consumption surveys food supply surveys, 1104 household food use data, 1104–1105 household inventories, 1104 individual consumption studies, 1105 models calendar model, 1107–1108 contact factors for dietary exposure model, 1103–1104 general dietary model, 1102–1103 general exposure model, 1101–1102 probabilistic models, 1103 screening and probabilistic models, 1103 overview, 1099–1101 quality audit and validation, 1112 residue estimation in foods, 1106 software for assessment Calendex, 1112–1113 Cumulative and Aggregate Residue Evaluation System, 1113 Customized Dietary Assessment Software, 1114 Dietary Exposure Evaluation Model, 1113–1114 Dietary Exposure Potential Model, 1114 Dietary Record Generator, 1114 REx, 1115 SHEDS-Wood, 1115 toxicity data for exposure context, 1108 uncertainty on exposure assessment, 1110 frequency, 7–8 human study measurements dermal exposure absorbent samplers, 1278–1279 air concentration, 1278 conventions, 1280 dyes, 1280 washing, 1279–1280 differential protection technique, 1281–1282 oral exposure, 1280–1281 respiratory exposure air concentration, 1276

2327

trapping toxicants in inhaled air, 1277–1278 Smyth technique, 1281 models comprehensive exposure models, 1002 human model types, 998–1002 overview, 997–998 pesticide regulation support chlorpyrifos exposure, 1013–1016 chromated copper arsenate exposures to children, 1013 screening-level exposure models advective transport, 1000 ambient environment, 1001 complexity and level of detail, 1000–1001 diffusive transport, 999–1000 fugacity models, 999 indoor environment, 1002 mass balance, 999 source-to-dose exposure modeling human activity-based exposure modeling, 1006–1008 reconstruction and inverse modeling, 1010–1011 sensitivity and uncertainty analysis, 1011–1012 steps, 1002–1005 toxicokinetic modeling, 1008–1010 nonoccupational exposure assessment ancillary information collection, 986 dermal exposure direct methods, 984 indirect methods, 984–985 diet exposure direct methods, 985–986 indirect methods, 985 inhalation exposure, 983–984 pesticide measurements general principles, 987–988 performance requirements, 989–991 techniques, 988–989, 991 sample collection biomonitoring samples, 986 criteria for collection method selection, 981, 983 multimedia collection methods, 982–983 soil, 986 water, 986 study design, 978–981 residential, see Residential exposure assessment routes in toxicity testing, 61–62 science-to-outcome framework, 972 terminology, 973, 996–997 turf pesticides, see Turf pesticides Exposure mitigation measures (EMMs) basics of process, 1150–1151 California, 1159–1160, 1166–1169 Environmental Protection Agency use, 1159–1160 fieldworkers, 1163–1165 initiation, 1151 overview, 1149–1150 pesticide handlers and users, 1162–1163

practicalities and limitations, 1160 protection factors, 1159–1162 residents and bystanders, 1165–1166 Extoxnet, 1364 Eye irrigation, poisoning management, 1299

F

FAO, see Food and Agriculture Organization Fat body fat pesticide storage, 925 toxicity testing influences of dietary fat, 74 Fate, environmental, see Environmental transport and fate Federal Food, Drug, and Cosmetic Act (FFDCA) aggregate exposure, 1375 consumer right to know, 1376 Delaney clause, 1374 estrogenic substances screening program, 1376 historical perspective, 1371 infants and children provisions, 1375–1376 safe pesticide definition, 1374–1375 tolerance reassessment, 1376–1377 toxicity mechanism and community risk assessment, 1375 Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA) advisory panel, 1373 antimicrobial pesticides, 1374 data collection, 1374 emergency suspension, 1373 minor-use pesticides, 1373–1374 overview, 1349–1350, 1371–1372 reduced-risk pesticides, 1374 registration of pesticides, 1377–1378 registration renewal, 1373, 1378–1379 risk assessment guidelines, 349–350 special review process, 1379–1380 tolerance reevaluation, 1373 toxicological database, 337–338 Fenarimol, skin testing, 759 Fenoxycarb developmental and reproductive toxicity, 414–415 skin testing, 740 Fenpropathrin metabolism, 1645–1646 properties, 1645 Fenthion age differences in toxicity, 825 developmental and reproductive toxicity, 411 Fenvalerate metabolism, 1646–1648 properties, 1646 skin testing, 746–747 FFDCA, see Federal Food, Drug, and Cosmetic Act FFP, see Fresh frozen plasma FIFRA, see Federal Insecticide, Fungicide, and Rodenticide Act Fipronil age differences in toxicity, 829–830 GABAA receptor modulation, 809 veterinary toxicology, 295

Index

2328

Flavin monooxygenase (FMO) biotransformation, 867–869, 894, 898, 906 organophosphorous insecticide metabolism, 1401–1402 Florasulam properties, 1871 structure, 1865 toxicity animals, 1871–1873 humans, 1873 uses, 1871 Fluazifop-butyl, skin testing, 777 Fluazolate, see Protoporphyrinogen oxidase-inhibiting herbicides Flucythrinate metabolism, 1648–1649 properties, 1648 Flufenpyr-ethyl, see Protoporphyrinogen oxidase-inhibiting herbicides Flumethrin metabolism, 1649 properties, 1649 Flumetsulam properties, 1873 skin testing, 777 structure, 1865 toxicity animals, 1873–1875 humans, 1875 uses, 1873 Flumiclorac, see Protoporphyrinogen oxidase-inhibiting herbicides Flumioxazin, see Protoporphyrinogen oxidase-inhibiting herbicides Fluometuron properties, 1728 structure, 1727 toxicity studies in animals, 1728 uses, 1728 Fluoroacetamide animal toxicity, 2161–2162 human toxicity, 2162–2163 poisoning treatment animals, 2162 humans, 2163 properties, 2160 structure, 2154 toxicokinetics, 261 uses, 2160–2161 Fluoroethanol animal toxicity, 2163 human toxicity, 2163 properties, 2163 structure, 2163 Fluoroglycofen, see Protoporphyrinogen oxidase-inhibiting herbicides Flusilazole, skin testing, 760 Fluthiacet-methyl, see Protoporphyrinogen oxidase-inhibiting herbicides Fluvalinate metabolism, 1649–1650 properties, 1649 FMO, see Flavin monooxygenase FOCUS models, exposure estimation, 1205 Folpet developmental and reproductive toxicity, 415

history and use, 1016 human risk assessment, 1939–1940 metabolism, 1920–1922 overview, 1915–1916 physical properties and chemical reactions, 1917–1920 skin testing, 754 toxicology acute toxicity, 1922–1924 carcinogenicity, 1933–1937 chronic toxicity, 1926–1927 developmental toxicity, 1927–1928 mechanism of toxicity, 1937–1938 mutagenicity, 1928–1933 overview, 1916–1917 reproductive toxicity, 1928 subchronic toxicity, 1924–1926 Fomesafen, see Protoporphyrinogen oxidaseinhibiting herbicides Food and Agriculture Organization (FAO), poisoning surveillance, 1348 Food Quality Protection Act (FQPA) child health and regulation of pesticides, 821–822 developmental and reproductive toxicity testing impact, 390 exposure assessment requirements, 974–975 safety factor for endosulfan, 515–516 Fosetyl-aluminum, skin testing, 760 Fosthiazate, skin testing, 737 FQPA, see Food Quality Protection Act Fresh frozen plasma (FFP), poisoning management, 1302 Freund’s complete adjuvant test, skin testing, 702 Fumigants, see Soil fumigants Fungicides biosynthesis targeting, 109–110 growth regulator targeting, 114–115 market, 106 nonspecific targets, 112–115 primary targets, 103–104 resistance, 105 respiration targeting, 112 secondary targets, 104 skin testing, 719–725, 753–764 turf, 1049–1054

G

GABAA receptor cyclodiene resistance, 2083 modulation dieldrin, 807–808 fipronil, 809 hexachlorocyclohexane, 806–808 mutations and cyclodiene resistance, 809 pyrethroids, 805, 1680–1682 picrotoxinin binding, 2078–2079 polychlorocycloalkane interactions, 2082–2088 topography, 2082–2083 Gas chromatography, pesticide residue analysis with gas chromatography–mass spectrometry, 1120–1121 Gastric lavage, poisoning management, 1298

Gene–pesticide interactions carcinogenesis, 526 environmental exposure assessment, 528–529 gene polymorphisms and pesticide detoxification butyrylcholinesterase, 539 carboxyl esterases, 538–539 cytochromes P450, 529–536 dopamine transporter, 540–541 glutathione S-transferase, 539–540 NAD(P)H:quinone oxidoreductase, 1, 541 P-glycoprotein, 541 paraoxonase, 536–537 Parkinson’s disease, 526–528 prospects for study, 541–542 Genotoxicity testing, see also specific pesticides bacterial reverse mutation assay, 360–361 cytogenetic assay in vivo, 361–362 human biomonitoring, 369–373 mammalian cell mutation assay, 361 micronucleus assay, 362 overview, 357–358 pesticide findings 1,4-dichlorobenzene, 369 bentazon, 367 chlorpyrifos, 367 endosulfan, 367 ethylene dibromide, 365 ethylene oxide, 362, 365 hexachlorobenzene, 366–367 methyl bromide, 365–366 o-phenylphenol, 367–368 propoxur, 366 table, 363–364 requirements in United States, 359, 361 risk assessment, 373–374 short-term tests, 358–360 Gliocladium virens, microbial pest control and toxicity, 455 Glue boards, vertebrate pest control, 282 Glufosinate animal toxicity acute and chronic, 185–186 mutagenicity, oncogenicity, and reproduction, 187 symptoms and pharmacology, 185–187 formulations and uses, 184 human poisoning and treatment, 187–188 mechanism of action, 184 metabolism and excretion, 184–185 physicochemical properties, 184 Glutamate receptors, modulation by pyrethroids, 806 Glutathione (GSH), liver levels in bromobenzene toxicity, 44 Glutathione S-transferase (GST) detoxification reactions, 872–873 gene polymorphisms and pesticide susceptibility, 539–540 proteomics analysis of pesticide induction, 610 Glyophosate animal toxicity acute toxicity, 1968 carcinogenicity, 1969 developmental and reproductive toxicity, 1969–1970

Index

dose studies, 1968 genotoxicity, 1969 sensitization, 1968 toxicokinetics, 1968–1969 history and uses, 1968 human toxicity, 1970–1971 overview, 1967 properties, 1967–1968 skin testing, 771–772 structure, 1967 Gossyplure, see Pheromones Graded responses, measurement, 33 GSH, see Glutathione GST, see Glutathione S-transferase Guinea pig maximization test, skin testing, 702

H

Haber’s rule, logtime–logdosage curve, 31–32 Halosulfuron-methyl, skin testing, 776–777 Hazard quotient, ecotoxicological risk assessment, 1207 Hazen equation, 1208 HCH, see Hexachlorocyclohexane HCNB, see Hexachloronorbornenes HDMR, see High dimensional model representation Hemoglobin, phosphine effects, 2262 Heptachlor, overview, 2071–2073 Herbicides, see also Ethylenebisdithiocarbamates; Imidazolinone herbicides; Protoporphyrinogen oxidase-inhibiting herbicides; Symmetrical triazine herbicides specific herbicides biosynthesis targeting, 108–109 growth regulator targeting, 114–115 market, 106 nonspecific targets, 112–115 photosynthesis and pigment targeting, 107–108 primary targets, 105–106 respiration targeting, 112 skin testing, 727–733, 767–781 Hexachlorobenzene developmental and reproductive toxicity, 415, 419 genotoxicity testing, 366–367 Hexachlorocyclohexane (HCH), GABAA receptor modulation, 806–808 Hexachloronorbornenes (HCNB), structure– toxicity relationships, 2076–2078 High dimensional model representation (HDMR), sensitivity and uncertainty analysis, 1012 High-performance liquid chromatography (HPLC) log P determination, 1231–1232 pesticide residue analysis with liquid chromatography–mass spectrometry, 1121–1123 HPLC, see High-performance liquid chromatography Human studies absorption measurement carbaryl, 1284 carcinogens, 1284 DDT, 1283–1284

2329

difficulties, 1283 herbicides, 1284 overview, 1282–1283 parathion, 1284 animal data limitations, 1287–1288 case studies of poisonings, 1255–1256 closing ceremony, 1270 codes, 1258–1259 court action, 1263–1264 design dose selection, 1266 double-blind studies, 1269–1270 initiation, 1269 parameter selection, 1264–1266 volunteer selection, 1266–1267 workers as volunteers, 1267–1269 Environmental Protection Agency risk assessment and data utility, 1285–1290 examples of pesticide studies, 1271–1274 exposure measurement dermal exposure absorbent samplers, 1278–1279 air concentration, 1278 conventions, 1280 dyes, 1280 washing, 1279–1280 differential protection technique, 1281–1282 oral exposure, 1280–1281 respiratory exposure air concentration, 1276 trapping toxicants in inhaled air, 1277–1278 Smyth technique, 1281 group review, 1260–1261 guidelines, 1259 informed consent, 1264 insurance, 1270 investigator precautions, 1270–1271 legislation France, 1262 Sweden, 1263 United Kingdom, 1262–1263 United States, 1262 moral obligations, 1259–1260 motivation of volunteers, 1271 species-to-species extrapolations, 1288 types of studies, 1257–1258 use experience, 1256 Humidity percutaneous absorption effects, 884 toxicity testing influences, 82 Hydramethylnon, skin testing, 751 Hydrogen cyanide, injury recovery, 1 Hydrolases, biotransformation, 870 Hydrophobicity, pesticides log P determination direct partitioning methods, 2129–1231 high-performance liquid chromatography, 1231–1232 nonexperimental estimation additive nature, 1232–1233 computer-aided procedures, 1237–1239 empirical procedures, 1233–1237 potentiometric titration, 1231

physicochemical significance aquatic toxicity, 1245–1248 bioaccumulation, 1243–1245 soil behavior, 1239, 1241–1242 2-(Hydroxymethylamino)-ethanol, skin testing, 734–735 Hypersensitivity, see also Photocontact dermatitis dosage–response relationships, 50

I

ICD, see Irritant contact dermatitis Ideal gas law, dose times concentration comparison, 4, 8 IFCS, see Intergovernmental Forum on Chemical Safety IGRs, see Insect growth regulators Imazalil developmental and reproductive toxicity, 420 skin testing, 760 Imazamethabenz-methyl, see Imidazolinone herbicides Imazamox, see Imidazolinone herbicides Imazapic, see Imidazolinone herbicides Imazapyr, see Imidazolinone herbicides Imazaquin, see Imidazolinone herbicides Imazethapyr, see also Imidazolinone herbicides skin testing, 779 Imidacloprid acetylcholine receptor modulation, 809–811 acute toxicity, 2057–2058 carcinogenicity, 2059–2060 chemistry, 2055–2056 chronic toxicity, 2059–2060 developmental and reproductive toxicity, 422, 427, 2060 metabolism, 2057 mutagenicity, 2060 neurotoxicity, 2061–2063 nicotinic activity insects, 2056 mammals, 2056 reproductive toxicity, 2061 skin testing, 751 subchronic toxicity, 2058–2059 toxicokinetics, 2057 veterinary toxicology, 294 Imidazolinone herbicides absorption, distribution, metabolism, and excretion, 1861–1862 animal toxicity acute toxicity, 1854–1858 chronic toxicity, 1859–1861 genotoxicity, 1863 oncogenicity, 1860–1861 pathology, 1862–1863 reproductive toxicity, 1862 subchronic toxicity, 1855, 1859 poisoning treatment, 1863 properties, 1853 types and structures, 1853, 1855 use experience, 1863 uses, 1854 Imiprothrin metabolism, 1650–1651 properties, 1650

Index

2330

Immunoassay, pesticide residue analysis, 1123 Immunotoxicity carbamates animal studies nonrodents, 486 rodents, 485 human studies, 486 immune system overview, 483–484 organochlorines animal studies nonrodents, 487–488 rodents, 486–487 human studies, 488 organophosphates animal studies nonrodents, 489 rodents, 488–489 human studies, 489 phenoxy compounds, 489–490 pyrethrins animal studies nonrodents, 490–491 rodents, 490 human studies, 491 regulation guidelines, 492 testing, 492–493 triazenes animal studies nonrodents, 491–492 rodents, 491 human studies, 492 Incremental repeated acquisition task, National Center for Toxicological Research Operant Test Battery, 853 Individual differences, toxicity testing, 70 Indoor residual spraying (IRS), vector-borne disease prevention, 248–249 Informed consent, human studies, 1258–1259, 1264 Injury factors affecting chronicity, 10–11 nature, 10 reversibility, 11–12 Insect growth regulators (IGRs) developmental and reproductive toxicity diflubenzuron, 414 fenoxycarb, 414–415 methoprene, 414 veterinary toxicology, 295 Insecticide biosynthesis targeting, 109–110 environmental and health concerns, 258–259 growth regulator targeting, 112–113 market, 106 neurotoxicity, 106 nonresidual insecticides, 260–261 nonspecific targets, 112–115 primary targets, 103, 105 residual insecticides, see Residual insecticides respiration targeting, 112 skin testing, 711–719, 735–753 Insecticide-treated bed nets (ITN), vector-borne disease prevention, 249–250 Integrated pest management (IPM)

advisory services, 309–311 avoidance practices, 308–309 biological controls and pesticide interactions, 309 continuum, 305–306 cultural and physical suppression, 308 field scouting decision support, 307 monitoring, 306–307 overview, 231, 238–239, 303–305 pesticide resistance, 306 pheromone use, 307–308 prevention practices, 308 reduced-risk pesticides, 307–308 turf, 1055, 1069 vector management in vector-borne disease mosquito abatement districts, 239 non-insecticidal methods, 240–241 pesticides acaricides, 244–245 classification, 241–244 ectoparasites, 244 indoor residual spraying, 248–249 insecticide-treated bed nets, 249–250 mosquito insecticides, 245–247 personal protection with repellents, 247–248 Intelligence quotient (IQ), National Center for Toxicological Research Operant Test Battery correlation, 854 Interaction, compounds antagonism, 58 laboratory test influences, 60–61 measurement, 56–58 mechanisms dynamics, 60 enzyme induction, 59–60 kinetics, 60 potentiation, 58–59 systematic study, 58 Intergovernmental Forum on Chemical Safety (IFCS), poisoning surveillance, 1348–1349 Iodomethane, see Soil fumigants IPM, see Integrated pest management Iprodione, skin testing, 760–761 IQ, see Intelligence quotient Irritant contact dermatitis (ICD) agricultural workers acute dermatitis, 653 cumulative dermatitis, 653–655 pesticide irritant types, 653–655 plants as irritants, 655–657 clinical patterns acute, 650 airborne, 651 cumulative, 651 delayed-acute, 650 exsiccation eczematoid, 651 friction dermatitis, 651 irritant reaction, 650 pustular, 651 suberythematous irritation, 651 subjective/sensory irritation, 650–651 traumatic, 651 definition, 647

diagnosis, 651–652 factors influencing irritant potential chemical factors, 647 endogenous patient characteristics, 648 physical factors, 647–648 testing bioengineering approaches for irritant identification, 649–650 chamber scarification test, 649 cumulative irritation testing, 649 immersion tests, 649 novel approaches, 650 patch testing, 649 IRS, see Indoor residual spraying Isobenzan, overview, 2071–2073 Isodrin, overview, 2066–2071 Isolation, toxicity testing influences, 74–75 Isoproturon properties, 1729 structure, 1728 toxicity studies in animals, 1729–1730 uses, 1729 Isoxaben, skin testing, 779 ITN, see Insecticide-treated bed nets Ivermectin, see also Avermectins age differences in toxicity, 829 exposure in humans, 2101–2102

J

Jasmolin, see Pyrethrins Joint probability curve, ecotoxicological risk assessment, 1209–1210

K

Kadethrin metabolism, 1652 properties, 1651–1652 Kairomones, see Pheromones Kasugamycin animal toxicity, 177 formulations and uses, 176 human toxicity, 177 mechanism of action, 176–177 metabolism and excretion, 177 physicochemical properties, 176 stability, 176 structure, 174 Kepone, see Chlordecone Ketoconazole developmental and reproductive toxicity, 420 steroidogenesis inhibition, 555–556 thyroid hormone disruption, 560 Kidney biotransformation, 908–909 glomerular filtration, 962 paraquat effects, 1787 tubular reabsorption, 962 tubular secretion, 962–963

L

Lactation DDT secretion in milk, 2003 toxicity susceptibility, 70 Lactofen, see Protoporphyrinogen oxidaseinhibiting herbicides

Index

Lagendium giganteum, microbial pest control and toxicity, 455–456 LC 50, dispersed toxicants, 33 LD, 50 90-dose LD, 50 comparison with one-dose test, 18 determination, 19 one-dose LD, 50 confidence limits and reproducibility, 16 corresponding ED values, 15–16 curve shape, 13–15 determination, 16–17 dose volume, 17 overview, 12–13 Lighting, toxicity testing influences, 79 D-Limonene, flea control, 293 Linalol, flea control, 293 Lindane developmental and reproductive toxicity, 409 GABAA receptor interactions, 2085 immunotoxicity, 486 overview, 2066–2071 skin testing, 748 Linuron androgen receptor function effects, 554 developmental and reproductive toxicity, 399 Lipid peroxidation paraquat effects, 1785–1786 phosphine effects, 2262 Lipoproteins, cyprodinil response, 1908–1909 Liver, ethylenebisdithiocarbamate effects, 1698 Livestock protection collar, coyote control, 282–283 LLNA, see Local lymph node assay Local lymph node assay (LLNA), skin testing, 702–703 Log P, see Hydrophobicity, pesticides Logtime–logdosage curve curve shape, 29–30 delayed toxicity, 28–29 Haber’s rule, 31–32 predictive value, 30–31 LT 50, principles, 27–28 Lung biotransformation, 908 excretion, 963 ozone toxicity, see Ozone paraquat effects accumulation, 1781–1782, 1792 efflux, 1782 fibrosis prevention, 1793, 1809–1810 pathology, 1776–1777

M

M-44, coyote control, 283 Magnesium phosphide, skin testing, 764 Magnetic resonance imaging, animal study endpoint, 850–851 Malaoxon acute toxicity, 1540 formation, 1527 long-term toxicity, 1541 short-term exposure, 1540–1541 Malathion

2331

absorption, distribution, metabolism, and excretion, 1531–1532 acute toxicity, 1532–1533 biotransformation, 1403, 1405 biotransformation, 1529–1530 carcinogenicity, 1536 chronic toxicity, 1535–1536 developmental and reproductive toxicity, 411–412, 1536–1537 exposure, 1531 genotoxicity, 1534–1535 human volunteer studies, 1538–1540 immunotoxicity, 488–489 impurities, 1528–1529 mechanism of action, 1528 neurotoxicity, 1537–1538 overview of use, 1527–1528 physicochemical properties, 1528 short-term toxicity dermal, 1533–1534 inhalation, 1534 oral, 1533 skin testing, 737 toxicology study review, 1532 Mancozeb, see also Ethylenebisdithiocarbamates immunotoxicity, 486 skin testing, 755–756 Maneb, see also Ethylenebisdithiocarbamates developmental and reproductive toxicity, 420 skin testing, 755–756 Margin of exposure (MOE), definition, 7 Mass spectrometry (MS) metabolomics, 631–633, 635 pesticide analysis gas chromatography–mass spectrometry, 1120–1121 liquid chromatography–mass spectrometry, 1121–1123 overview, 988, 991 proteomics isotope labeling, 607 liquid chromatography–mass spectrometry, 605–606 tandem mass spectrometry, 607 MBTC, see Methylene bis(thiocyanate) MCPA, see 2-Methyl-4-chlorophenoxy acid MCPP, see 2-(2-Methyl-4-chlorophenoxy) propionic acid MDR1, see P-glycoprotein Mean residence time (MRT), pharmacokinetics, 927 Medical use, pesticides, 1256, 1544–1545 Melaleuca oil, flea control, 293 Merphos, skin testing, 772 Metabolism, see Biotransformation Metabolomics applications, 636–638 data mining, 634–635 mass spectrometry, 631–633, 635 nuclear magnetic resonance, 633–635 overview, 627–629 sample pretreatment and extraction, 630–631

Metaldehyde, veterinary toxicology, 296 Metam-sodium acute toxicity, 2294 adjustment factors, 575 chronic toxicity, 2298, 2300–2302 developmental and reproductive toxicity, 420–421, 2297–2298 genotoxicity, 2295–2297 neurotoxicity, 2302–2303 overview, 2293–2294 skin testing, 767 subchronic toxicity, 2294–2295 uses, 2293 Metarhizium anisopliae, microbial pest control and toxicity, 454–455 Methamidophos, skin testing, 737–738 Methanearsonic acid, skin testing, 779 Methidathion, skin testing, 738 Methiocarb, skin testing, 740 Methomyl, skin testing, 741 Methoprene, developmental and reproductive toxicity, 414 Methoxychlor animal toxicity mechanisms, 2011 organ and tissue findings, 2011–2013 reproductive toxicity, 2013–2014 toxicokinetics, 2010–2011 developmental and reproductive toxicity, 408–409 estrogen receptor function effects, 553–554 human toxicity, 2014 immunotoxicity, 486–487 metabolism, 899 properties, 2010 skin testing, 748 structure, 1976 uses, 2010 Methyl bromide, see also Soil fumigants acute exposure risk assessment bystander exposure, 346–349 exposure estimates, 346 toxicology, 344–346 acute toxicity, 2268–2269 adjustment factors, 574 buffer zones, 348–349 chronic toxicity, 2273–2274 developmental toxicity, 427, 2272–2273 exposure mitigation measures, 1167–1169 genotoxicity testing, 365–366, 2270–2272 human exposure, 2277 metabolism absorption, 2275–2276 distribution, 2276 excretion, 2277 metabolites, 2276–2277 neurotoxicity, 2274–2275 oncogenicity, 2273–2274 reproductive toxicity, 427, 2272–2273 skin testing, 765–766 soil applications, 315, 322–323 subchronic toxicity, 2269–2270 uses, 2267–2268

Index

2332

N-Methyl carbamate insecticides see also specific insecticides cumulative risk assessment needs, 1591, 1593 relative potency factor approach, 1593–1594 physiologically-based pharmacokinetic modeling calibration and validation, 1595–1597 carbofuran, 1598–1599 cumulative risk assessment, 1601–1602 exposure assessment as input, 1600–1601 exposure-related dose estimating model, 1595 multiple chemicals, 1599–1600 parameters, 1594–1595 postmodel construction analysis, 1597–1598 prospects, 1603 types and structures, 1592 2-Methyl-4-chlorophenoxy acid (MCPA) properties, 1830 skin testing, 773 2-(2-Methyl-4-chlorophenoxy) propionic acid (MCPP, properties, 1830 Methylene bis(thiocyanate) (MBTC), skin testing, 761 Methyl iodide crop production, 2311 environmental fate soil distribution, 2314–2316 emissions, 2313–2314 leaching, 2316 mechanism of transformation, 2313 overview, 2312 temperature and moisture effects, 2312–2313 water, 2311–2312 exposure quantification microbe control, 2310–2311 nematode control, 2310 weed control, 2309–2310 overview, 2307 properties, 2307–2308 registration, 2308–2309 Methyl isothiocyanate, skin testing, 766–767 2-Methyl-4-isothiazolin-3-one, skin testing, 706–707 Methylisothiocyanate, see Soil fumigants Methyl parathion, acute exposure risk assessment bystander exposure, 342–344 occupational exposure, 341–342 toxicology, 340–341 Methyl thiophanate, developmental and reproductive toxicity, 421 Metiram, see Ethylenebisdithiocarbamates Metofluthrin metabolism, 1652 properties, 1652 Metolachlor carcinogenicity mechanisms, 1767–1768 properties, 1762 skin testing, 769

structure, 1754 toxicity in animals, 1762–1763 uses, 1762 Metosulam properties, 1875 structure, 1865 toxicity animals, 1875–1877 humans, 1877 uses, 1875 Metribuzin, skin testing, 779 Microbial pest control agents (MPCAs) California utilization trends by type, 442–444 market, 441 toxicity Bacillus cereus, 451 Bacillus sphaericus, 451–453 Bacillus thuringiensis, 447–451 baculovirus, 456 Burkholderia cepacia, 453–454 Gliocladium virens, 455 Lagendium giganteum, 455–456 Metarhizium anisopliae, 454–455 rabbit hemorrhagic disease virus, 456–457 testing requirements overview, 445 Tier I, 445–446 Tier II, 446 Tier III, 446–447 Micronucleus assay, genotoxicity testing, 362 Mildiomycin animal toxicity, 178 formulations and uses, 178 mechanism of action, 178 physicochemical properties, 178 stability, 178 structure, 174 Milinate developmental and reproductive toxicity, 398 skin testing, 774 steroidogenesis inhibition, 556–557 Mirex, overview, 2075–2076 Mitigation, see Exposure mitigation measures MOE, see Margin of exposure Mosquito, see Vector-borne disease MPCAs, see Microbial pest control agents MRT, see Mean residence time MS, see Mass spectrometry Multiple Scenario Risk Assessment Tool (MUSCRAT), 1205 MUSCRAT, see Multiple Scenario Risk Assessment Tool Mustard seed, rodenticide use, 2206 Mutagenesis, dosage–response relationships, 50 Myopathy, organophosphorous insecticide poisoning, 1554–1555

N

Nabam, see Ethylenebisdithiocarbamates NAD(P)H:quinone oxidoreductase 1, gene polymorphisms and pesticide susceptibility, 541 Naled, skin testing, 738 -Naphthylthiourea (ANTU) animal toxicity, 2168–2171 human toxicity

dosage response, 2171 incidents, 2171 use experience, 2171 overview, 2167–2168 poisoning treatment animals, 2170–2171 humans, 2171 properties, 2168 structure, 2164 toxicokinetics, 2168 uses, 2168 National Agricultural Workers Survey (NAWS) data sources, 1346 description, 1346 findings, 1346–1347 period of data collection, 1346 report frequency, 1346 National Center for Health Statistics (NCHS) case definition for acute pesticide-related illness and injury, 1353–1354 hospital discharge survey, 1344–1345 multiple cause of death data, 1342–1343 National Center for Toxicological Research Operant Test Battery, nonhuman primate studies, 852–855 National Institute for Occupational Safety and Health (NIOSH), 1364 National Pesticide Information Center (NPIC), 1350, 1365 National Poison Data System (NPDS), 1314–1317 National Research Council (NRC), Environmental Protection Agency guidance on risk assessment, 1288–1290 NAWS, see National Agricultural Workers Survey NCHS, see National Center for Health Statistics Neem tree, see Azadirachtin Neonicotinoid insecticides imidacloprid, see Imidacloprid overview, 2055–2056 Nerium indicum, rodenticide use, 2206 Neuropathy target esterase (NTE) biochemical properties, 1443 biomarker and biosensor applications, 1442–1443 catalytic domain, 1445, 1447 enzymology, 1443–1444 identification, 1437 isolation and tissue distribution, 1444 molecular biology, 1445–1446 organophosphorous pesticide-induced delayed neuropathy activity testing, 1495–1497 human risk assessment, 1440–1442 mechanisms loss of function in phospholipid metabolism, 1448–1451 toxic gain of function, 1447–1448 role, 1437–1438, 1490–1492 structure–activity relationships, 1439–1440 organophosphorous insecticide reactivity, 1435–1437

Index

overview, 1435 prospects for study, 1451 Neurotoxicity, see also Behavior specific pesticides; Organophosphorous pesticide-induced delayed neuropathy acetylcholine receptor modulation imidacloprid, 809–811 pyrethroids, 806 assessment acute toxicity testing guidelines, 591–592 behavioral changes, 587–589 electrophysiology, 589–590 neurochemical endpoints, 590 neuropathology, 591 calcium channel modulation by pyrethroids, 805, 1677–1679 chloride channel modulation by pyrethroids, 808, 1679–1680 chronic poisoning workup, 1308 dosage–response relationships, 47 GABAA receptor modulation dieldrin, 807–808 fipronil, 809 hexachlorocyclohexane, 806–808 mutations and cyclodiene resistance, 809 pyrethroids, 805, 1680–1682 glutamate receptor modulation by pyrethroids, 806 insecticides, 106 regulatory studies cholinesterase comparative sensitivity studies, 599 organophosphorous compound delayed neurotoxicity, 592–595 peripheral nerve function, 597–598 schedule-controlled operant behavior, 596–597 sensory evoked potentials, 598–599 sodium channel modulation by DDT and pyrethroids gating kinetics, 799–800 open channel properties, 801 pyrethroids amplification of toxicity, 802–803 electrophysiology, 1673–1674 factors affecting toxicity, 803–804 gating kinetics, 799–800 mutations in resistance, 806–807 open channel properties, 801 sensitivity of different channels, 801–802, 1674–1675 sites of action, 801, 1675–1676 state dependency of channel modification, 800–801 state-dependent actions, 1675 temperature dependence of toxicity, 803 toxicity correlation, 1677 vitamin E alleviation of paresthesia, 804–805 state dependency of channel modification, 800–801 systems biology approach developmental toxicity, 794 literature review, 794–796 overview, 793–794

2333

Nicotine absorption, metabolism, and excretion, 129 agricultural uses, 128 animal toxicity acute, 130–132 poisoning incidents, 132 teratogenic, carcinogenic, and mutagenic effects, 130, 132 treatment, 132 formulations, 128 human toxicity pathology, 134 poisoning incidents, 133 reproduction, 132–133 treatment, 133–134 mechanism of action, 129 neonicotinoid insecticides, see Imidacloprid pharmacology, 129–130 physicochemical properties, 128 stability, 128 structure, 127 Nitenpyram, veterinary toxicology, 294 Nitrofen, developmental and reproductive toxicity, 399 Nitrogen oxides, ozone reactivity, 571 NMR, see Nuclear magnetic resonance No-observed-effect level, 37 Nonhuman primates, see Animal models Norbormide animal toxicity, 2202–2203 human toxicity, 2203–2204 properties, 2202 structure, 2200 uses, 2202 Norflurazon, skin testing, 779–780 NPDS, see National Poison Data System NPIC, see National Pesticide Information Center NRC, see National Research Council NTE, see Neuropathy target esterase Nuclear magnetic resonance (NMR), metabolomics, 633–635 Nutrition, toxicity testing influences dietary fat effects, 74 dietary protein effects, 73–74 general nutritional condition, 72 starvation effects, 72–73

O

Obidoxiume chloride, pesticide poisoning management, 1560 Occupational exposure see also specific pesticides assessment fluorescent tracers and visible dyes, 1131 hand and head exposure, 1132–1134 inhalation exposure, 1134–1135 patch method, 1130–1131 study design for agricultural workers, 1129–1130 test subjects, 1130 whole body method, 131–1132 biological monitoring of absorbed dose, 1135 databases

Agricultural Reentry Task Force transfer coefficient database, 1150–1153 European Union, 1146–1149 Pesticide Handlers Exposure Database, 1142–1146 re-entry exposure database, 1149, 1151–1153 endosulfan, see Endosulfan guidelines for assessment, 1128 mitigation, see Exposure mitigation measures modeling Agricultural Handlers Exposure Database, 1149 generic modeling, 1139–1140 German BBA model, 1146–1149 Predictive Operator Exposure Model, 1147–1149 tiered approach, 1141–1142 organophosphorous insecticide biomonitoring blood cholinesterases, 1575–1577 hypersusceptible subject detection, 1577 urinary alkylphosphates, 1575 routes dermal, 1127 inhalation, 1127–1128 oral, 1128 sulfur combustion episodes, 1899 eye, 1897 respiratory tract, 1897 skin, 1895–1897 systemic illness, 1897–1899 Octhilinone, skin testing, 706–707 N-Octyl-bicycloheptene dicarboximide, skin testing, 742 Oncogenicity, see Carcinogenicity OPIDN, see Organophosphorous pesticide-induced delayed neuropathy Organophosphorous pesticide-induced delayed neuropathy (OPIDN) clinical manifestations animals, 1481–1482 general features, 1567–1568 humans, 1481 laboratory findings, 1568–1569 differential diagnosis, 1569 esterase studies, 1438 etiology, 1566–1567 hen test, 1439 history of study, 1479–1480 neuropathology in animals, 1482–1490, 1569 neuropathy target esterase activity testing, 1495–1497 human risk assessment, 1440–1442 mechanisms loss of function in phospholipid metabolism, 1448–1451 toxic gain of function, 1447–1448 role, 1437–1438, 1490–1492 structure–activity relationships, 1439–1440 pathogenesis, 1490–1493, 1567 regulatory studies, 592–595 susceptibility factors, 1493–1495 treatment, 1569

Index

2334

Organophosphorous insecticides see also specific pesticides biotransformation conjugations, 1404 hydrolysis, 1402–1404 oxidations, 1400–1402 reductions, 1402 chemistry and biochemistry, 1545–1547 classification and nomenclature, 1395–1396, 1533 clinical manifestations of poisoning cardiac manifestations, 1549 central nervous system manifestations, 1549–1550 differential diagnosis, 1556 general features, 1547–1548 laboratory findings, 1550–1552 late complications, 1561–1564 respiratory failure, 1548–1549 short-term complications, 1552–1556 drugs, 1544–1545 historical perspective, 1395 intermediate syndrome clinical manifestations, 1565 etiology, 1564 pathogenesis, 1564 long-term exposure effects central nervous system, 1570 miscellaneous effects, 1574–1575 neurobehavioral effects, 1573–1574 peripheral nervous system, 1571–1573 psychiatric effects, 1573 neuropathy, see Organophosphorous pesticide-induced delayed neuropathy occupational exposure biomonitoring blood cholinesterases, 1575–1577 hypersusceptible subject detection, 1577 urinary alkylphosphates, 1575 pharmacokinetics compartmental models, 1411–1412 disposition and clearance, 1415–1417 exposure monitoring and quantification, 1417–1423 health risk assessment children, 1426–1429 insecticide mixtures, 1424–1426 overview, 1409–1411 physiologically-based pharmacokinetic models, 1412–1416, 1430 reactions, 1398 structures and chemistry, 1480–1481 synthesis, 1396–1398 targets see Acetylcholinesterase; Butyrylcholinesterase; Neuropathy target esterase treatment of poisoning antidotes, 1472–1473, 1557–1560 further exposure minimization, 1556–1557 sequence of treatment, 1561 supportive treatment, 1561 warfare agents, 1544 Oryzalin, skin testing, 770 Oxadiargyl, see Protoporphyrinogen oxidase-inhibiting herbicides

Oxadiazon, see also Protoporphyrinogen oxidase-inhibiting herbicides skin testing, 780 Oxamyl, skin testing, 741 Oxydemeton-methyl, skin testing, 738 Oxyfluorfen, see also Protoporphyrinogen oxidase-inhibiting herbicides skin testing, 780 Oxythioquinox, skin testing, 752 Ozone formation, 571 nitrogen oxide reactivity, 571 toxicity animal studies asthma exacerbation, 579 epithelial permeability, 577 lung development, 578–579 lung host defense, 578 lung structure and cellularity changes, 578 human studies asthma induction and exacerbation, 580 hospital admissions, 580–581 inflammation, 579–580 mortality, 581 volatile organic compounds and formation, 571

P

Pancreatitis, organophosphorous insecticide poisoning, 1553–1554 Paraoxon, biotransformation, 1402 Paraoxonase (PON), gene polymorphisms and pesticide susceptibility, 536–537 Paraquat animal studies absorption, 1777–1778 acute toxicity, 1772–1774 carcinogenicity, 1775–1776 central nervous system effects, 1788–1791 distribution, 1778–1779 excretion, 1780–1781 kidney effects, 1787 lung effects accumulation, 1781–1782 efflux, 1782 pathology, 1776–1777 metabolism, 1779–1780 mutagenicity, 1775–1776 poisoning treatment, 1791–1793 reproductive toxicity, 1776 sensitization, 1774 signs of toxicity, 1772 subchronic toxicity, 1774–1775 developmental and reproductive toxicity, 398 human toxicity absorption, 1805 distribution, 1805–1806 excretion, 1806 experimental exposure, 1793 metabolism, 1806 pathology, 1806–1807 poisoning atypical cases, 1798–1800

fulminant or hyperacute toxicity, 1801–1804 laboratory findings, 1804–1805 mild or subacute toxicity, 1801 moderate to severe toxicity, 1801 overview, 1794–1795 treatment, 1807–1811 use experience, 1795–1798 properties, 1771 skin testing, 768 structure, 1771 synonyms, 1771 toxicity mechanisms lipid peroxidation, 1785–1786 mitochondria role, 1786–1787 NADPH oxidation, 1786 oxidative stress, 1783–1785, 1787 uses, 1771–1772 veterinary toxicology, 296 Parathion absorption measurements in humans, 1284 age differences in toxicity, 823, 825 behavior effects in development, 839–841 biotransformation, 1402 developmental and reproductive toxicity, 411 percutaneous absorption, 683–684 skin testing, 738–739 systemic absorption, 683, 685 Parkinson’s disease, gene–pesticide interactions, 526–528, 540–541 Passive transport, absorption, 878–879 Patch test, see Skin testing PBO, see Piperonyl butoxide PBPK models, see Physiologically-based pharmacokinetic models PCCP, see Poison control center PDR, see Potential dose rate PDR, see Potential dose rate PEM, see Polioencephalomalacia Pendimethalin, skin testing, 770–771 Pennroyal oil, flea control, 293–294 Penoxsulam properties, 1878 structure, 1865 toxicity animals, 1878–1879 humans, 1879 uses, 1878 Pentachloronitrobenzene, skin testing, 761–762 Pentachlorophenol, developmental and reproductive toxicity, 421 Percutaneous absorption anatomical site differences, 687–689, 881–882 assays absolute topical bioavailability, 683 biologic response, 686 penetration in vivo versus in vitro, 686–687 radioactivity in excreta, 683–685 skin flaps, 685 stripping method, 685–686 biomonitoring, 696 chemicals in clothing, 689–692 factors affecting humidity, 884

Index

soil, 884–886 temperature, 887 modeling biosensors, 696 cluster analysis, 692, 694 dermal assessment estimate method, 695 dermal exposure ranking method, 694–695 estimation and assessment exposure, 695–696 Risk Assessment of Occupational Dermal Exposure, 696 parathion, 683–684 pesticide formulation and mixture effects, 882–883 skin decontamination, 696–698 Permethrin age differences in toxicity, 828 metabolism, 1653–1654 properties, 1653 Pesticide Handlers Exposure Database (PHED) acute exposure risk assessment, 339–340 grading criteria, 1143 guidance for use, 1143–1144 historical perspective, 1142–1143 limitations, 1146 overview, 1143 surrogate exposure scenarios, 1145 Pets Environmental Protection Agency exposure methodology activity patterns, 1087 formulation properties, 1086 overview, 1084–1086 postapplication dermal and incidental ingestion exposure variables, 1087 research needs, 1087–1089 ownership demographics, 1078 pesticide interactions, 879 spot-on treatment application, 1079–1080 applicator and postapplicator exposure assessment, 1084–1084 exposure monitoring studies, 1079–1084 veterinary products and services, 1078–1079 P-glycoprotein (MDR1) gene polymorphisms and pesticide susceptibility, 541 overview, 2104–2106 placenta, 2106–2107 protein binding of drugs, 879–880 Pharmacokinetics compartmental models multicompartment models intravenous bolus dose, 932–933 extravascular dose, 933–934 one-compartment model extravascular dose, 930–931 intravenous bolus dose, 928–930 storage with repeated extravascular exposure, 931–932 noncompartmental models, 926–927 physiologically-based models, 934–937 PHED, see Pesticide Handlers Exposure Database

2335

Phenothrin metabolism, 1654–1655 properties, 1654 skin testing, 744 o-Phenylphenol genotoxicity testing, 367–368 skin testing, 734 Phenylurea herbicides see Diuron; Fluometuron; Isoproturon Pheromones environmental fate, 169 examples codlemone, 172 disarture, 171 gossyplure, 171–172 kairomones curcurbitacins, 172–173 trimedlure, 172 (Z)-9-tricosene, 171 integrated pest management, 307–308 mechanism of action, 169 overview, 167–168 stability, 169 toxicity, 169–171, 173 uses and formulations, 168–169 Phorate, biotransformation, 1402 Phosmet, skin testing, 739 Phosphamidon, immunotoxicity, 489 Phosphine acute toxicity, 2260–2261 biochemical effects, 2261–2263 chemistry, 2259 dose response, 2261 genotoxicity, 2263 poisoning treatment, 2263–2264 properties, 2259 regulation, 2264 reproductive toxicity, 2263 skin testing, 764 sources, 2260 toxicokinetics, 2261 Photocontact dermatitis clinical workup, 664–665 histology, 666 management, 665–667 photoallergic contact dermatitis, 662–664 phototesting, 665–666 phytophotodermatitis clinical presentation, 661–662 hyperpigmentation, 662 plant induction, 662 ultraviolet radiation and photosensitivity, 661 Photoperiodicity, chemical toxicity testing effects, 80 Photosensitization, chemical toxicity testing influences, 80–81 Photosynthesis, herbicide targets, 107–108 Physiologically-based pharmacokinetic (PBPK) models N-methyl carbamate insecticides calibration and validation, 1595–1597 carbofuran, 1598–1599 cumulative risk assessment, 1601–1602 exposure assessment as input, 1600–1601

exposure-related dose estimating model, 1595 multiple chemicals, 1599–1600 parameters, 1594–1595 postmodel construction analysis, 1597–1598 prospects, 1603 organophosphorous insecticides, 1412–1416, 1430 overview, 934–937, 1008–1010 Phytophotodermatitis (PPD) clinical presentation, 661–662 hyperpigmentation, 662 plant induction, 662 Picardin acute toxicity, 2223 chemistry, 2219 chronic toxicity, 2224–2225 dermal absorption, 2226–2227 developmental toxicity, 225 effectiveness against insect vectors, 2220–2221 genotoxicity, 2225 mechanism of action, 2219–220 metabolism, 2221–2223 neurotoxicity, 2226 oncogenicity, 2225 overview, 2219 reproductive toxicity, 2225–2226 safety evaluation, 2227–2228 subchronic toxicity, 2224 toxicokinetics, 2221–2223 Picloram, skin testing, 773–774 Pigment synthesis, herbicide mechanisms, 107–109 Piondrel, skin testing, 753 Piperonyl butoxide (PBO) acute toxicity, 2128 chemistry and formulations, 2127 chronic toxicity, 2136–2140 developmental toxicity, 2134–2136 exposure assessment, 2145–2146 genotoxicity, 2140–2143 human toxicity and experience, 2143–2144 pharmacodynamics, 2144–2145 reproductive toxicity, 2131–2134 risk characterization cancer, 2146–2147 dietary risk, 2147 non-dietary risk, 2147 skin testing, 742 subchronic toxicity, 2128–2131 uses, 2127–2128 Placenta, pesticide permeability, 925 Plasma clearance (Cl), pharmacokinetics, 927 POEM, see Predictive Operator Exposure Model Poison control center (PCCP) description, 1314–1317 National Poison Data System, 1314–1317 pesticide exposures by type, 1317–1323 report frequency, 1317 syndromic surveillance, 1323–1324 target population, 1317

Index

2336

Poisoning acute poisoning epidemiology, 1299 symptoms and signs, 1299–1302 treatment, 1302 chronic poisoning assessment exposure history, 1305–1306 signs, 1306–1308 symptoms, 1306 background accumulation, 1304 carcinogenesis risks, 1308–1309 neurotoxicity workup, 1308 overview, 1302–1303 symptoms and signs, 1305 toxicity types, 1303–1304 treatment, 1309 management see also specific pesticides activated charcoal, 1297–1298 cathartics, 1298–1299 eye irrigation, 1299 gastric lavage, 1298 overview, 1296–1297 skin decontamination, 1297 whole-bowel irrigation, 1299 pesticide types and toxicity, 1295–1296 surveillance, see Surveillance Polioencephalomalacia (PEM), sulfur induction, 1893 PON, see Paraoxonase Potential dose rate (PDR), calculation, 1043–1044 Potentiometric titration, log P determination, 1231 PPD, see Phytophotodermatitis Pralidoxamine, pesticide poisoning management, 1302, 1558–1559 Prallethrin metabolism, 1656 properties, 1655 Predictive Operator Exposure Model (POEM), 1147–1149 Pregnancy, toxicity susceptibility, 70 Pressure, toxicity testing influences, 79 Probabilistic risk assessment, see Ecotoxicological risk assessment Prochloraz, steroidogenesis inhibition, 557 Procymidone, androgen receptor function effects, 555 Prodione, androgen receptor function effects, 555 Progressive ratio task, National Center for Toxicological Research Operant Test Battery, 852 Prometon, see also Symmetrical triazine herbicides skin testing, 775 Prometryn, see also Symmetrical triazine herbicides skin testing, 775 Propachlor carcinogenicity mechanisms, 1767–1768 properties, 1763 skin testing, 769 structure, 1754 toxicity

animals, 1763–1765 humans, 1765 uses, 1763 Propanil immunotoxicity, 487 skin testing, 780–781 Propargite, skin testing, 752 Propargyl bromide, skin testing, 766 Propazine, see also Symmetrical triazine herbicides Propioconazole, developmental and reproductive toxicity, 420 Propoxur genotoxicity testing, 366 immunotoxicity, 485 skin testing, 741 Prostaglandin synthetase, biotransformation, 869–870 Protection factors, exposure mitigation measures, 1159–1162 Proteomics bacterial pesticide degradation studies catabolic and metabolic networks, 618 cell membrane proteins, 618–619 overview, 611–618 stress response and adaptation, 619 bioinformatics, 607–610 glutathione S-transferase induction by pesticides, 610 isotope labeling, 607 liquid chromatography–mass spectrometry, 605–606 overview, 603 tandem mass spectrometry, 607 two-dimensional gel electrophoresis, 604–605 Protoporphyrinogen oxidase-inhibiting herbicides diphenyl ether protoporphyrinogen oxidaseinhibiting herbicides, 1733 environmental impact degradation, 1742 ecotoxicology, 1742–1745 soil interactions, 1742 mammalian toxicology metabolism, 1746–1747 mutagenicity, 1745 oral, 1743, 1745 porphyrin metabolism effects, 1745–1746 skin, 1743, 1745 teratogenicity, 1745 non-diphenyl ether protoporphyrinogen oxidase-inhibiting herbicides, 1733–1736 overview, 1733 plant behavior absorption and metabolism, 1737–1738 mechanism of action, 1738–1740 resistance genetic engineering, 1742 modes, 1740–1741 uses application mode, 1736–1737 crops and weeds, 1734, 1736 Public health surveillance, see Surveillance PubMed, 1365

Pyraflufen-ethyl, see Protoporphyrinogen oxidase-inhibiting herbicides Pyrethrins, see also Pyrethroids animal toxicity acute, 123–125 chronic and subchronic, 123, 125 endocrine effects, 126 irritation and inhalation, 125–126 teratogenic, carcinogenic, and mutagenic effects, 126 treatment, 126 developmental and reproductive toxicity, 413–414 formulation and uses, 121–122 human toxicity irritation and sensitization, 127 systemic poisoning, 126–127 treatment, 127 hypothalamic–pituitary–gonadal axis disruption, 559–560 immunotoxicity animal studies nonrodents, 490–491 rodents, 490 human studies, 491 mechanism of action, 122–123 metabolism, 1657 metabolism and excretion, 123 overview, 120, 1656 physical and chemical properties, 1656–1657 skin testing, 741–742 stability of extracts, 121 structures, 120, 1665 types, 120–121, 1656 veterinary toxicology, 292–293 Pyrethroids, see also specific Pyrethroids; Pyrethrins acetylcholine receptor modulation, 806 acute toxicity, 1668–1669 age differences in sensitivity, 1670 calcium channel modulation, 805, 1677–1679 chloride channel modulation, 808, 1679–1680 development, 1666 GABAA receptor modulation, 805, 1680–1682 glutamate receptor modulation, 806 intoxication neurochemical consequences, 1670 syndromes, 1669–1670 mechanism of action, 1667 neurotoxicity behavioral neurotoxicity effects acoustic startle response, 1671–1672 conditioned behavior, 1672 motor activity, 1671 regulatory guidelines, 1672 dermal exposure, 1672–1673 developmental neurotoxicity, 1673 human reports, 1670–1671 overview, 1635, 1665 sodium channel modulation and pyrethroids amplification of toxicity, 802–803 electrophysiology, 1673–1674

Index

factors affecting toxicity, 803–804 gating kinetics, 799–800 mutations in resistance, 806–807 open channel properties, 801 sensitivity of different channels, 801–802, 1674–1675 sites of action, 801, 1675–1676 state dependency of channel modification, 800–801 state-dependent actions, 1675 temperature dependence of toxicity, 803 toxicity correlation, 1677 vitamin E alleviation of paresthesia, 804–805 structure–activity relationships, 1666–1667 structure–toxicity relationships, 1669 Pyriminil animal toxicity, 2164–2165 human toxicity dosage response, 2166 incidents, 2165–2166 laboratory findings, 2166–2167 pathology, 2167 poisoning treatment animals, 2165 humans, 2167 properties, 2164 structure, 2164 toxicokinetics, 2165 uses, 2164 Pyroxsulam properties, 1880 structure, 1865 toxicity animals, 1880–1881 humans, 1881 uses, 1880

R

Rabbit hemorrhagic disease virus, microbial pest control and toxicity, 456–457 Radiation chemical toxicity testing influences, 79 dosage response studies, 21 Recovery, definition, 4 Red blood cell, chronic toxicity testing, 470 Red squill, see Scilliroside Registration of pesticides, see Federal Insecticide, Fungicide, and Rodenticide Act Relative potency factor (RPF), N-methyl carbamate insecticide cumulative risk assessment, 1593–1594 Reproductive toxicity see also specific pesticides epidemiology, 383–384 exposure adult reproductive toxicants, 385 prenatal reproductive toxicants, 384–385 prepubertal reproductive toxicants, 385 timing, 384 fungicides benomyl, 415 dinocap, 415 ethyl dibromide, 419 folpet, 415

2337

hexachlorobenzene, 415, 419 imazalil, 420 ketoconazole, 420 maneb, 420 metam sodium, 420–421 methyl thiophanate, 421 pentachlorophenol, 421 propioconazole, 420 table, 416–419 terrazole, 421 vinclozolin, 421–422 herbicides amitrole, 397 bromoxynil, 397 2,4-D, 391, 397 dinoseb, 398 diquat, 398 diuron, 399 ethofumesate, 398 ethyl dipropyl thiocarbamate, 398 linuron, 399 milinate, 398 nitrofen, 399 paraquat, 398 table, 392–397 triazines, 399 insect growth regulators diflubenzuron, 414 fenoxycarb, 414–415 methoprene, 414 insecticides aldrin, 407 amitraz, 407 bendiocarb, 412 carbaryl, 413 carbofuran, 412 chlordecone, 409 chlorpyrifos, 410–411 DDT, 408 diazinon, 411 dibromochloropropane, 407–408 dimethoate, 411 fenthion, 411 lindane, 409 malathion, 411–412 methoxyxchlor, 408–409 overview, 399, 407 parathion, 411 pyrethrins, 413–414 table, 400–407 thiodicarb, 413 thiram, 413 trichlorfon, 412 zineb, 413 mechanisms of pesticide action, 385 methyl bromide, 427 overview, 383 rodenticides, 422 testing dose selection, 388 exposure assessment, 389–390 Food Quality Protection Act impact, 390 interpretation, 388–389 principles, 386–388 regulatory history, 386

species selection, 388 statistical evaluation, 389 veterinary pesticides, 422–427 Residential exposure assessment literature review, 1093–1094 overview, 1091–1093 regulatory guidance and data sources, 1094 research needs, 1094–1095 Residual insecticides crack and crevice treatments, 262–263 general treatments, 261–262 spot treatments, 262 Residue analysis development of method cleanup, 1120 extraction, 1119–1120 multiresidue methods, 1118–1119 separation and quantification capillary electrophoresis, 1123–1124 gas chromatography–mass spectrometry, 1120–1121 immunoassay, 1123 liquid chromatography–mass spectrometry, 1121–1123 validation of method confirmation, 1117–1118 independent laboratory validation, 1118 radiovalidation, 1118 Resistance, overview, 107, 911–912 Resmethrin metabolism, 1657–1658 properties, 1657 skin testing, 744 REx, dietary exposure assessment, 1115 Ricin animal toxicity acute, 197–198 treatment, 198–199 distribution, metabolism, and excretion, 197 human poisoning and treatment, 199–200 mechanism of action, 196–197 overview, 196 physicochemical properties, 196 stability, 196 uses, 196 Risk assessment, see Acute exposure; Carcinogenicity testing; Ecotoxicological risk assessment Risk Assessment of Occupational Dermal Exposure (RISKOFDERM), percutaneous absorption modeling, 696 Risk communication audience, 1174, 1176 benefits information, 1185–1186 ecotoxicological risk assessment, 1212–1213 media role, 1174 newspaper headline examples, 1175 perceptions of risk who believes, 1179–1181 why people believe, 1176–1179 presentation of risk concepts, 1181–1185 strategy rationale, 1173–1174 RISKOFDERM, see Risk Assessment of Occupational Dermal Exposure

Index

2338

Risk perception intuitive toxicology, 1388–1390 overview of studies, 1381–1384 risk communication who believes, 1179–1181 why people believe, 1176–1179 social, cultural, and political influences on, 1384–1388 Rodenticides, see also Bromethalin; Chloralose; Fluoroacetamide; Fluoroethanol; -Naphthylthiourea; Norbormide; Pyriminil; Ricin; Salmonella; Scilliroside; Sodium fluoroacetate; Strychnine anticoagulant rodenticides brodifacoum, 2184–2188 bromadiolone, 2192–2193 chlorophacinone, 2188–2190 coumafuryl, 2182 difenacoum, 2190–2192 difethialone, 2193–2194 diphacinone, 2182–2184 overview, 2171–2173 resistance, 2173–2174 warfarin, 2174–2182 baits overview, 277–278 preparation, 278, 280 rodenticides, 278–279 banned compounds, 2205 developmental and reproductive toxicity, 422 overview, 2153–2154 prospects, 2205–2206 veterinary toxicology, 296 vitamin D-related compounds cholecalciferol, 2197–2199 ergocalciferol, 2194–2197 Rotenone animal toxicity acute, 139–140 chronic, 141 irritation, 140 pathology, 141 poisoning symptoms, 140–141 reproduction, 141 treatment, 141 cytotoxicity and mutagenicity, 138–139 formulations, 136 human poisoning, 141–142 mechanism of action, 136 metabolism and excretion, 136–137 pharmacology, 137–138 physicochemical properties, 135 stability, 135–136 structure, 135 RPF, see Relative potency factor Ryanodine animal toxicity, 147–148 formulations and uses, 146 mechanism of action, 146 metabolism and excretion, 147 pharmacology, 146–147 physicochemical properties, 146 stability, 146 structure, 145

Ryanodol animal toxicity, 147–148 formulations and uses, 146 mechanism of action, 146 metabolism and excretion, 147 pharmacology, 146–147 physicochemical properties, 146 stability, 146 structure, 145

S

Sabadilla alkaloids, see Cevadine; Veratridine Safe residue level (SRL), calculation, 1043–1044 Salmonella human poisoning incidents, 200 rodenticide use, 200 Saltidin, see Picardin Scilliroside animal toxicity, 194–195 formulations and uses, 189 human poisoning and treatment, 194–196 mechanism of action, 193–194 physicochemical properties, 193 structure, 188 Season, toxicity testing influences, 82 Sensor Event Notification System for Occupational Risk (SENSOR), 1326–1328 SENSOR, see Sensor Event Notification System for Occupational Risk Sensory evoked potentials, neurotoxicity assessment, 598–599 Sethoxydim, skin testing, 781 Sevin, see Carbaryl Sex differences, toxicity testing, 70 SHEDS-Wood, dietary exposure assessment, 1115 Silvex, properties, 1830 Simazine, see also Symmetrical triazine herbicides skin testing, 775 Skin absorption, see Percutaneous absorption decontamination, 1297 Skin testing see also specific pesticides adjuvants, 781 antimicrobial agents and disinfectants, 706–710, 734–735 fumigants, 725–727, 764–767 fungicides, 719–725, 753–764 herbicides, 727–733, 767–781 insecticides and repellents, 711–719, 735–753 protocols and regulation dermal irritation tests, 701–701 illness surveillance data integration, 704–706 quantitative structure–activity relationships, 703–704 regulatory decisions, 704 sensitization testing, 702–704 skin reaction patterns, 701 Smyth technique, exposure assessment, 1281 Sodium borate, skin testing, 750

Sodium channel, modulation by DDT and pyrethroids gating kinetics, 799–800 open channel properties, 801 pyrethroids amplification of toxicity, 802–803 electrophysiology, 1673–1674 factors affecting toxicity, 803–804 gating kinetics, 799–800 mutations in resistance, 806–807 open channel properties, 801 sensitivity of different channels, 801–802, 1674–1675 sites of action, 801, 1675–1676 state dependency of channel modification, 800–801, 1675 temperature dependence of toxicity, 803 toxicity correlation, 1677 vitamin E alleviation of paresthesia, 804–805 state dependency of channel modification, 800–801 Sodium dichloroisoccyanurate dihydrate, 707 Sodium fluoroacetate animal toxicity, 2154–2157 human toxicity dosage response, 2159–2160 incidents, 2157–2159 laboratory findings, 2160 pathology, 2160 use experience, 2159 poisoning treatment animals, 2157 humans, 2160 properties, 2154 structure, 2154 toxicokinetics, 2156 uses, 2154 Sodium methyldithiocarbamate, immunotoxicity, 485 Sodium tetrathiocarbonate, see also Soil fumigants adjustment factors, 575 Soil hydrophobicity of pesticides and behavior, 1239, 1241–1242 sample collection for nonoccupational exposure assessment, 986 Soil fumigants adjustment factors for fumigants chloropicrin, 575 dazomet, 575 1,3-dichloropropene, 574–575 metam sodium, 575 methyl bromide, 574 sodium tetrathiocarbonate, 575 adsorption, 320–321 applications chloropicrin, 324 1,3-dichloropropene, 323–324 dimethyl disulfide, 325 iodomethane, 323 methyl bromide, 315, 322–323 methylisothiocyanate, 324–325 sodium tetrathiocarbonate, 325–326

Index

degradation, 320 distribution and efficacy assessment, 321–322 emission minimization application methods, 326 chemical amendment, 326–327 irrigation, 326 mass balance, 327 organic amendment, 327 plastic films, 326 target area treatment, 327 environmental concerns, 319 fumigation methods chemigation, 318 emission minimization, 326–327 high pressure liquid gas injection, 318 low pressure liquid chisel injection, 318 power mulch or rototiller, 318 historical perspective, 315 physicochemical properties, 316–318 skin testing, 725–727, 764–767 structures, 317 types, 315–317 volatilization, 319–320 Species differences, toxicity testing absorption dermal absorption, 84–85 gastrointestinal absorption, 84 biotransformation phase I, 85–86 phase II, 86 distribution, 85 excretion urinary versus biliary, 86–87 urinary versus fecal, 87 humans versus animals, 67–69 parasites and hosts, 65 vertebrates, 65–66, 69–70 Spinosad animal toxicity acute, 165–166 reproduction, carcinogenicity, and mutagenicity, 167 formulations and uses, 163 human toxicity, 167 mechanism of action, 164 metabolism and excretion, 164–165 overview, 162 physicochemical properties, 163 spinosyn structures, 163 stability, 163–164 Spinosyns, see Spinosad SRL, see Safe residue level Starvation, toxicity testing influences, 72–73 Statistics control group abnormal values, 47 developmental and reproductive toxicity testing, 389 dosage level selection, 35–36 geometric mean, 45–46 reproducibility, 46–47 small dosage effects beneficial effects of small dosages, 39–40 chemical basis of thresholds in dosage– response relationships, 43–44

2339

logprobit model and quantitative analysis, 40–43 lowest effect level, 37 no-observed-effect level, 37 overview, 36–37 threshold levels as biological facts, 37–39 subjects number number of independent units, 34 identity of sampling unit, 34 randomization, 35 Strain differences, toxicity testing, 70 Streptomycin animal toxicity acute and chronic, 181 pharmacology, 181–182 treatment, 182 formulations and uses, 180–181 human toxicity, 182 mechanism of action, 181 metabolism and excretion, 181 physicochemical properties, 180 stability, 180 structure, 180 Strychnine absorption, metabolism, and excretion, 189–190 animal toxicity acute, 190–191 pathology, 191 treatment, 191–192 formulations and uses, 189 human toxicity acute, 192 metabolism and excretion, 192 pathology, 192 poisoning and treatment, 192–193 mechanism of action, 189 physicochemical properties, 189 structure, 188 Subjects number identity of sampling unit, 34 number of independent units, 34 randomization, 35 Sulfentrazone, see Protoporphyrinogen oxidase-inhibiting herbicides Sulfluramid, skin testing, 752–753 Sulfometuron, skin testing, 777 Sulfosate, skin testing, 771 Sulfur acute toxicity dermal, 1890–1891 inhalation, 1891 oral, 1890 environmental fate, 1889–1890 human health effects, 1894–1895 irritation eye, 1891 skin, 1892 occupational exposure combustion episodes, 1899 eye, 1897 respiratory tract, 1897 skin, 1895–1897 systemic illness, 1897–1899

pesticide use, 1889 sensitization, 1892 skin testing, 762–763 veterinary effects, 1893–1894 Sulfur dioxide fumigant use, 1892 skin testing, 762–763 toxicology, 1892–1893 Sulfuryl fluoride acute toxicity, 2246–2247 chemistry, 2245 chronic toxicity dog, 2253–2254 mouse, 2252–2253 rat, 2253 genotoxicity, 2255 human toxicology, 2256–2257 metabolism, 2255–2266 poisoning treatment, 2247–2248 repeated exposure studies, 2248 reproductive toxicity, 2245–2255 risk characterization, 2257 subchronic toxicity dog, 2251–2253 history of study, 2247 mouse, 2248–2249 rabbit, 2251 rat, 2249–2251 teratotoxicity, 2254 uses, 2245 Surveillance Bureau of Labor Statistics case definition, 1339 data sources, 1339 description, 1339 findings, 1341–1342 period of data collection, 1339 report frequency, 1339 target population, 1339 Colorado State University studies of hospitalized poisoning cases, 1343–1344 Department of Pesticide Regulation, California case definition, 1333–1334 data sources, 1334 description, 1332–1333 findings, 1334–1338 period of data collection, 1334 report frequency, 1334 target population, 1334 Environmental Protection Agency reporting requirements, 1350 epidemiology, see Epidemiologic studies evaluation of systems, 1351–1352 Food and Agriculture Organization, 1348 information evaluation, 1363–1364 Intergovernmental Forum on Chemical Safety, 1348–1349 limitations case definitions, 1356–1357 denominators in rate calculation, 1354–1356 sensitivity, 1357 National Agricultural Workers Survey data sources, 1346

Index

2340

Surveillance (Continued) description, 1346 findings, 1346–1347 period of data collection, 1346 report frequency, 1346 National Center for Health Statistics case definition for acute pesticide-related illness and injury, 1353–1354 hospital discharge survey, 1344–1345 multiple cause of death data, 1342–1343 overview, 1313–1314 pesticide information resources, 1364–1365 poison control centers description, 1314–1317 National Poison Data System, 1314–1317 pesticide exposures by type, 1317–1323 report frequency, 1317 syndromic surveillance, 1323–1324 target population, 1317 public health uses of data, 1357–1358 South Carolina hospital discharge surveys, 1345–1346 state-based surveillance systems advantages, 1331–1332 case definition, 1327 data sources, 1327–1328 findings, 1329–1331 overview, 1325–1327 period of data collection, 1328 report frequency, 1328 target population, 1328 strengthening of acute case surveillance, 1358–1360 types, 1313 World Health Organization, 1347–1348 Symmetrical triazine herbicides acute studies, 1713 carcinogenicity assessment, 1716–1717 classification of cancer, 1720–1721 epidemiology, 1719–1720 rat tumor formation mechanisms, 1717–1719 developmental and reproductive toxicity, 1714, 1716–1717 hazard identification, 1711–1714, 1721 mutagenicity, 1716 overview, 1711 repeat exposure and toxicity, 1713–1714 types and structures, 1711–1712 uses, 1711 Systems biology, see Neurotoxicity

T

2,4,5-T, see 2,4,5-Trichlorophenoxyacetic acid TCDD, see 2,3,7,8-Tetrachlorodibenzo-p-dioxin TCMTB, see Thiocyanomethylthiobenzothiazole TDE, see 1,1-Dichloro-2,2-bis (4-chlorophenyl)ethane

Temperature percutaneous absorption effects, 887 toxicity testing influences, 77–78 Temporal response differentiation, National Center for Toxicological Research Operant Test Battery, 852–853 TEPP, see Tetraethylpyrophosphate Teratogenesis, dosage–response toxicity relationships, 47–48 Terbuthylazine, see also Symmetrical triazine herbicides Terbutryn, see also Symmetrical triazine herbicides hydrophobicity estimation, 1240 Termite baits, 265–266 soil treatments, 263–265 Terrazole, developmental and reproductive toxicity, 421 Tetrachlorvinphos biotransformation, 1406 skin testing, 739 Tetraethylpyrophosphate (TEPP), oral toxicity, 54 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) chemical action, 9–10 concentration times time concept, 5–6 half-life, 12 impurity in Agent Orange, 56 species differences in toxicity, 65 Tetramethrin metabolism, 1658–1659 properties, 1658 skin testing, 744–745 TGA, see Thermogravimetric analysis Thermogravimetric analysis (TGA), volatile organic compounds, 572–573 Thidiazuron, skin testing, 777 Thiobencarb, skin testing, 774 Thiocyanomethylthiobenzothiazole (TCMTB), skin testing, 763 Thiodicarb, developmental and reproductive toxicity, 413 Thiophanate methyl, skin testing, 755 Thiram developmental and reproductive toxicity, 413 skin testing, 755–756 Thyroid, ethylenebisdithiocarbamate effects, 1691, 1695, 1697–1698 Thyroid hormone, pesticide disruption 2,4-D, 561 ketoconazole, 560 triclosan, 561 Time concentration times time concept, 4–8 definition, 6 Tolerance, overview of mechanisms, 911 Tolylfluanid mechanism of toxicity, 1938–1939 structure, 1938 Total body water, distribution, 924 Toxaphene, overview, 2075–2076

Toxicity definition, 3–4 injury duration factors affecting chronicity, 10–11 reversibility, 11–12 nature, 10 Toxicodynamics concentration times time concept, 5 definition, 3–4 Toxicokinetics concentration times time concept, 4–8 definition, 3–4 Toxicology, definition, 10 Tralomethtin metabolism, 1659 properties, 1659 Transgenics, vector-borne disease control, 241 Transport, environmental, see Environmental transport and fate Triadimefon, skin testing, 763 Triazine herbicides, see Symmetrical triazine herbicides Triazolpyrimidine herbicides, see Cloransulammethyl; Diclosulam; Florasulam; Flumetsulam; Metosulam; Penoxsulam; Pyroxsulam Tribufos, skin testing, 772 Tributyltin compounds, skin testing, 734 Trichlorfon, developmental and reproductive toxicity, 412 1,1-(2,2,2-Trichloroethylidene)-bis (4-chlorobenzene) (DDT) absorption measurements in humans, 1283–1284 age differences in toxicity, 71–72 animal toxicity absorption, 1981–1982 behavioral effects, 1992 carcinogenesis, 1989–1989 distribution, 1982 excretion, 1986 mechanism of action, 1986–1988 metabolism, 1982–1986 mortality causes, 1989 nervous system, 1988 pathology, 1992 poisoning treatment, 1992–1993 repeated dose response, 1979–1981 reproductive toxicity, 408, 1990–1992 symptomatology, 1977–1979 developmental and reproductive toxicity, 408 disease effects on toxicity, 76–77 enzyme induction, 50–51 formulations and production, 1976–1977 historical perspective, 333, 1975–1976 human toxicity cancer, 2000–2001 dose response, 2002 environmental exposure, 1998–2000 excretion of related compounds, 2003–2004 inhalation, 1995–1996 laboratory findings, 2004 milk secretion, 2003

Index

oral, 1993–1995 poisoning, 1996–1998 skin, 1995 storage, 2003–2003 therapeutic use, 1996 induction of microsomal enzymes, 942 metabolism and storage, 51–53, 69 nutritional effects on toxicity, 72–73 photodynamic action, 78 properties, 1976 resistance, 107 skin testing, 747 sodium channel modulation gating kinetics, 799–800 open channel properties, 801 state dependency of channel modification, 800–801 soil exposure, 885 species differences in toxicity, 65–66 structure, 1976 2,4,5-Trichlorophenoxyacetic acid (2,4,5-T), properties, 1830 Triclocarban, steroid enhancement, 557 Triclopyr, skin testing, 774 Triclosan steroid enhancement, 557 thyroid hormone disruption, 561 (Z)-9-Tricosene, see Pheromones Trigluralin, skin testing, 771 Trimedlure, see Pheromones Turf pesticides environmental issues drinking water quality, 1057–1058 groundwater, 1059–1063 surface water, 1058–1059 environment functional aspects, 1047–1048 usage, 1047 exposure activity patterns on turf effects, 1041–1042 assessment techniques, 1042–1043 biomonitoring and dosimetry, 1042, 1067 overview, 1037–1039 regulatory issues, 1043–1044 safe residue level calculation, 1043 transferrable residue assessment, 1039–1041 volatile and dislodgeable residues, 1063–1067 wildlife, 1068–1069 fungicides, 1049–1054 herbicides, 1055–1057 insecticides, 1050 management practices adjuvants, surfactants, and absorbents, 1071–1072 barriers for runoff containment, 1072 integrated pest management, 1055, 1069 irrigation, 1071 modeling and risk assessment, 1069–1071 market, 1048–1049 nematicides, 1050, 1055 Two-dimensional gel electrophoresis, proteomics, 604–605

2341

U

Ultra-low volume (ULV) techniques, mosquito control, 246–247 Ultraviolet radiation, chemical toxicity testing influences, 79–80 ULV techniques, see Ultra-low volume techniques Urticaria, see Contact urticaria syndrome

V

Validamycin A animal toxicity, 179–180 formulations and uses, 179 mechanism of action, 179 metabolism, 179 physicochemical properties, 178–179 stability, 179 structure, 174 Vector-borne disease arthropods human health impact, 237–238 types and pathogens, 232–235 enzootic cycles, 237 integrated pest management in vector management mosquito abatement districts, 239 non-insecticidal methods, 240–241 pesticides acaricides, 244–245 classification, 241–244 ectoparasites, 244 indoor residual spraying, 248–249 insecticide-treated bed nets, 249–250 mosquito insecticides, 245–247 personal protection with repellents, 247–248 mosquito pathogens, 236 surveillance, 236–237 transmission, 234–236 vectorial capacity, 236 Veratridine animal toxicity acute, 144 teratogenicity, 144 formulations and uses, 142–143 human toxicity irritation, 145 poisoning, 145 treatment, 145 mechanism of action, 143 metabolism and excretion, 144 pharmacology, 143–144 structure, 142 Vertebrate pests bird control flock dispersal agent, 283 poison bait, 283–284 contraceptives, 282 coyote control livestock protection collars, 282–283 M-44, 283 glue boards, 282 immobilizing agents, 277 lethal pesticides

fumigants, 281 overview, 275, 277 rodent baits, 277–281 tracking powders, 281–282 management restrictions, 272 nonlethal management, 272–273 overview, 271–272 problems, 272 repellents, 274–277 trapping, 273–274 Veterinary toxicology, see also Pets carcinogenesis, 291 developmental and reproductive toxicity, 422–427 human exposure, 291 intoxication acute, 290 chronic, 290–291 diagnosis, 296–297 frequency, 289–290 treatment, 297–298 pesticide categories amitraz, 295 cholinesterase inhibitors, 292 fipronil, 295 flea control, 293–294 insect growth regulators, 295 macrocyclic lactones, 294 metaldehyde, 296 neonicotinoids, 294–295 paraquat, 296 pyrethrins and pyrethroids, 292–293 rodenticides, 296 synergists and repellents, 295–296 pesticide use in livestock, 287 regulatory agencies in United States, 287–289 violative residues, 289 Vinclozolin androgen receptor function effects, 554–555 developmental and reproductive toxicity, 421–422 skin testing, 763–764 Vitamin E, alleviation of pyrethoid-induced paresthesia, 804–805 VOC, see Volatile organic compounds Volatile organic compounds (VOC) adjusting nonfumigants for application method effects on emissions, 575–576 adjustment factors for fumigants chloropicrin, 575 dazomet, 575 1,3-dichloropropene, 574–575 metam sodium, 575 methyl bromide, 574 sodium tetrathiocarbonate, 575 inventory calculations, 573 ozone interactions, 571 toxicity, see Ozone pesticide contributors, 572 reactivity, 576–577 special default emission potential, 574 thermogravimetric analysis, 572–573 Volume of distribution, 924

Index

2342

W

Warfarin animal toxicity, 2174–2177 human toxicity dosage response, 2181 experimental exposure, 2177–2178 incidents, 2180 laboratory findings, 2181–2182 pathology, 2182 therapeutic use, 2178–2180 use experience, 2180–2181 overview, 2171–2173 poisoning treatment animals, 2177 humans, 2182 properties, 2174 resistance, 2173–2174, 2176 structure, 2172

toxicokinetics, 2175–2176 uses, 2174 Water hydrophobicity and aquatic toxicity, 1245–1248 sample collection for nonoccupational exposure assessment, 986 toxicity testing influences chlorine content, 83 hardness, 82 pH, 82 turf pesticide environmental issues drinking water quality, 1057–1058 groundwater, 1059–1063 surface water, 1058–1059 WBI, see Whole-bowel irrigation White blood cell, chronic toxicity testing, 470 WHO, see World Health Organization

Whole-bowel irrigation (WBI), poisoning management, 1299 Worker exposure, see Occupational exposure Worker Protection Standard for Agricultural Pesticides (WPS), 1350–1351 World Health Organization (WHO), poisoning surveillance, 1347–1348 WPS, see Worker Protection Standard for Agricultural Pesticides

Z

Zinc borate, see Borates Zinc phosphide, skin testing, 764 Zineb, see also Ethylenebisdithiocarbamates developmental and reproductive toxicity, 413 skin testing, 755–756 Ziram, skin testing, 755–756